OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID



## SIDS Initial Assessment Report

After

SIAM 22

Paris, France, 18-21April 2006

1. Chemical Name: Ammonium Perfluor ooctanoate & Perfluorooctanoic Acid
2. CAS Number: 3825-26-1 & 335-67-1 3. Sponsor Country Environment and Exposure

Human Health part Germany Contact Point: BMU (Bundesministerium für Umwelt, Naturschutz und Reaktorsicherheit) Contact person: Alexander Nies Postfach 12 06 29 D- 53048 Bonn

US-EPA
4. Shared Partnership with: Environmental and Exposure part: UBA and BUA, Germany DuPont de Nemours (Belgium) BVBA 5. Roles/Responsibilities of the Partners:

• Name of industry sponsor /consortium DuPont de Nemours (Belgium) BVBA A Spinoystraat 6 B-2800 Mechelen Belgium Dr. Watze de Wolf Eric van Wely

• Process used Sharing process: US-EPA: Human Health part Germany: Environmental and Exposure part 6. Sponsorship History

• How was the chemical or category brought into the OECD HPV Chemicals Programme?

1OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

7. Review Process Prior to the SIAM: 1.) last literature search (update): June 2006 (Toxicology) June 2005 (Ecotoxicology): database s CA, biosis; search profile CAS-No. and special search terms. - March 2006 (Exposure): company information, databases as named above. 2.) Other review processes: Mammalian Toxicology: 1) Extern al Review of US EPA reportwritten procedure, April 2003; 2) U. S. EPA Science Advisory Board , February 2005 Ecotoxicology: OECD ICCA revi ew process including addition of few relevant current information and publications 8. Quality check process: Internal crosscheck 9. Date of Submission: 1.) January 20 th2006 for SIAM 22 2.) written procedure before SIAM 24 10. Comments: For the assessment of environmen tal hazards, with some minor exceptions, only data published before 2006, March1 st is considered.

2OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

SIDS INITIAL ASSESSMENT PROFILE

CAS No. 335-67-1 3825-26-1

Chemical Name Perfluorooctanoic Acid (PFOA) Ammonium Perfluorooctanoate (APFO)

Structural formula

COOH F FFFF F F

F FFFF F F F COOH•NH 3 F FFFF F F

F FFFF F F F

SUMMARY CONCLUSIONS OF THE SIAR Chemical IdentityFor the purposes of this document, the anion of PFO (perfluorooctanoate) is frequently referenced as PFOA or APFO. APFO and PFOA are so metimes used interchangeably. Perfluorooctanoic Acid (PFOA) is a fully fluorinated carbox ylic acid. APFO is the ammonium salt of PFOA.

Human Health
APFO is the ammonium salt of PFOA and the two substances are metabolically equivalent. PFOA is a strong acid and it is expected to dissociate in biological media. Several epidemiology and medical surveillance studies have been co nducted on workers employed at various APFO manufactur ing sites in the U.S. Most of the studies were cross-sectional and focused primarily on males. A retrospective cohort mortality study demonstrated a statistically significant association between prostate cancer mortality and employment duration in the chemical facility of a plant that manufactures PFOA. However, in an update to this study in which more specific exposure measures were use d, a significant association for prostate cancer was not observed. Other mortality studies lacked adequate exposure data which could be linked to health outcomes. A study which examined hormone levels in workers reported an increase in estradiol levels in workers with the highest PF OA serum levels; however, these results may have been confounded by body mass index. Cholesterol and triglyceride levels inworkers were positively associated with PFOA exposures, which is inconsistent with the hypolipidemic effects observed in rat studies. A statistically significan t positive association was reported for PFOA and T3 levels in workers but not for any other thyroid hormones. Little information is available concerning the pharmacokinetics of PFOA and its salts in humans. Preliminary results of a 5-year half-life study in 26 retired worker s indicate that the mean serum elimination half-life of PFOA in these wo rkers was 3.8 years (1378 days, 95% CI, 1131-1624 days) and the range was 1.5-9.1 years. The pharmacokinetics of PFOA in non-human prim ates has been studied both in classical intravenous elimination studies using three male and three female cynomolgus monkeys and in a sixmonth, repeat-oral-dose toxicology study in male cynomolgus monkeys with recovery. These studies confirmed urinary elimination as the primary excretion mode. From the intravenous study, in which males and females were given a 10 mg/kg dose, mean elimination half-lives were 20.9 days for males and 32.6 days for females. In the six-month study, monkeys were dosed with 3, 10, or 30/20 mg/kg-d ammonium PFOA. Steady-state serum concen tration was reached within

3OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

four to six weeks, with steady state levels lowe r than those that would be predicted based on the elimination rates and not in lin ear proportion to dose. Males in the six-month toxicology study had elimination half-life rates approximating 20 days.
Studies in adult rats have shown that the ammonium salt of PFOA (APFO) is absorbed following oral and inhalation exposure; le ss absorption occurs following dermal exposure. Serum pharmacokinetic parameters and the distribution of PFOA ha s been examined in the tissues of adult rats following administration by gavage and by intravenous (i.v.) and intraperitone al (i.p.) injection. PFOA distributes primarily to the liver, serum, and kidney, and to a lesser extent, other tissues of the body. It does not partition to th e lipid fraction or adipose tissu e. The distribution of PFOA ispredominantly extracellular. PFOA is not metabolized and there is evidence of enterohepatic circulation of the compound. The urine is the major r oute of excretion of PFOA in the female rat, while the urine and feces are both main routes of excretion in male rats. There are gender differences in the elimination of PFOA in adult rats following administration by gavage and by i.v. and i.p.injec tion. In female rats, following oral administration, estimates of the serum half-life were dependent on dose and ranged from approximately 2.8-16 hours, while in male rats estimates of the serum half-life fo llowing oral administrati on were independent of dose and ranged from approximately 138-202 hours. In female rats, elimination of PFOA appears to be biphasic with a fast phase and a sl ow phase. The rapid excret ion of PFOA by female rats is believed to be due to active renal tubular secretion (organic anion transporters ); this renal tubular secretion is believed to be hormonally controlled. Hormonal changes during pregnancy do not appear to cause a change in th e rate of elimination in rats.
Several recent studies have been conducted to examine the kineti cs of PFOA in the developing rat. These studies have shown that PFOA readily cr osses the placenta and is present in the breast milk of rats. The gender difference in elimina tion is developmentally regulated; between 4-5 weeks of age, elimination assumes the adult pa ttern and the gender difference becomes readily apparent. Distribution studies in the postweaning rat have shown that PFOA is distributed primarily to the serum, liver, and kidney. In acute toxicity studies in animals using APFO, the oral LD 50 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 inhalatio n exposure of 18.6 mg/l APFO for one hour in rats. The dermal LD 50 in rabbits was determined to be gr eater than 2000 mg/kg. APFO is a primary ocular irritant in rabbits, while the data regarding potential skin irritancy are conflicting.
APFO did not induce mutation in ei ther 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 c oncentrations. 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 ab errations were observed without metabolic activation. However, when tested with metabolic ac tivation, APFO induced significant increases in chromosomal aberrations and in polyploidy. APFO was negative in a cell transformation assay in mouse embryo fibroblasts and in the mouse micronucleus assay. Repeat-dose studies have been conducted in non-human primates. In a 13-week study with Rhesus monkeys, exposure to doses of 30 mg/kg-day or higher resulted in death. Clinical signs of toxicity were noted at doses as low as 3 mg/kg-day. Unlike rodent studies, analyses of the serum and liver levels did not reve al a gender difference in monkeys, but the sample size was very

4OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

small. In a 6-month study of male cynomolgus monkeys, there was a steep dose response curve for mortality. Increases in liver weight were no ted at doses as low as 3 mg/kg-day, but there was no evidence of peroxisome proliferato r-activated receptor alpha activity (PPAR α ). The LOAEL for the study was 3 mg/kg-day, and a NOAEL was not identified. Repeat-dose studies in rats and mi ce demonstrated that the liver is the primary target organ. Due to gender differences in elimination, adult male rats exhibit effects at lower administered doses than adult female rats. Dietary exposure to APFO fo r 90 days resulted in significant increases in liver weight and hepatocellular hypertrophy in fe male rats at 1000 ppm (76.5 mg/kg-day) and in male rats at doses as low as 100 ppm (5 mg/kg-da y). Chronic dietary exposure of rats to 300 ppm (males, 14.2 mg/kg-day; females, 16.1 mg/kg-day) APFO for 2 years resulte d in increased liver weight, hepatocellular hypertrophy, hematological e ffects, and testicular masses in males; and reductions in body weight and hema tological effects in females. The carcinogenic potential of PFOA has been inves tigated in two dietary carcinogenicity studies in rats. Under the conditions of these studies, ther e is some evidence that PFOA is carcinogenic, inducing liver tumors, Leydig cell tumors (LCT), and pancreatic acinar cell tumors (PACT) in male rats. The evidence for mammary fibroadenomas in the female rats is equivocal since the incidences were comparable to some historical background incidences. There is sufficient evidence to indicate that PFOA is a PPAR α -agonist and that the liver car cinogenicity (and toxicity) of PFOA is mediated by PPAR α in the liver. PFOA appears to be immunotoxic in mice. Feed ing 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 res ponse to horse red blood cells, and suppression of splenic lymphocyte proliferation. The suppre ssed 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 PPAR α was involved in causing the advers e effects to the immune system. In an oral prenatal developmental toxicity st udy in rats, the LOAEL and NOAEL for maternal toxicity were 150 mg/kg-day and 5 mg/kg-day, re spectively. There was no evidence of developmental toxicity after exposure to doses as high as 150 mg/kg-day. In a rat inhalation developmental toxicity study, the NOAEL and LOAEL for maternal toxicity were 1 and 10 mg/m3, respectively. The NOAEL and LOAEL for developmen tal toxicity were 10 and 25 mg/m3, respectively. In a rabbit oral prenatal developmental toxicity study there was a significant increase in skeletal variations after e xposure to 5 mg/kg-day APFO, an d the NOAEL was 1.5 mg/kg-day. There was no evidence of maternal toxicity at 50 mg/kg-day, the highest dose tested. In a mouse oral developmental toxicity study, there was evid ence of maternal toxicity and developmental toxicity and the authors calcula ted benchmark doses for a vari ety of endpoints. Decreased weight gain and increased liver weight was obs erved in adult females; the BMD5 and BMDL5 estimates for decreases in maternal weight ga in were 6.76 and 3.58 mg/kg, respectively, and the BMD5 and BMDL5 estimates for increases in ma ternal liver weight were 0.20 mg/kg and 0.17 mg/kg, respectively. The BMD5 and BMDL5 estimates for the incidence of fu ll-litter resorptions and neonatal mortality (determined by survival to weaning) observed at the 5 mg/kg/day dose group were 2.84 and 1.09 mg/kg, respectively. Signifi cant alterations in postnatal growth and development were observed at 1 and 3 mg/kg/day, with BMD5 and BMDL5 estimates of 1.07and 0.86 mg/kg, respectively, for decreased pup weight at weaning; and 2.64 and 2.10 mg/kg, respectively, for delays in eye opening. The BMD5 and BMDL5 estimates for reduced phalangeal ossification were <1 mg/kg. BMD5 and BMDL5 esti mates for reduced fetal weight at term were estimated to be 10.3 and 4.3 mg/kg, respectively. A variety of endpoints were eval uated throughout different lifesta ges in a two-generation reproductive toxicity study in rats exposed to 0, 1, 3, 10, or 30 mg/kg/day APFO. In that study, a re5OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

duction in F1 pup mean body weight on a litter ba sis was observed during lactation (sexes combined) in the 30 mg/kg-day gr oup. F1 male pups in the 10 and 30 mg/kg-day groups exhibited a significant reduction in body weight gain during days 8-50 po stweaning, and body weights were significantly reduced in the 10 mg/kg-day group beginning on postweaning day 36, and in the 30 mg/kg-day group beginning on postweaning day 8. F1 female pups in the 30 mg/kg-day group exhibited a significant reducti on in body weight gain on days 1-15 postweaning, and in body weights beginning on day 8 postweaning. Reproductive indices were not affected in the F1 animals. There was a significant increase in mortality mainly during the first few days after weaning, and a significant delay in the timing of sexual maturation for F1 male and female pups in the 30 mg/kg-day group. No effects were observed in th e F2 pups. 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. Adult systemic toxicity consisted of reductions in body weight in both the F0 and F1 animals.

Environment
Both PFOA and APFO are solid at environmenta l relevant temperatures. The melting point for PFOA is 54.3°C and the boiling point is 190°C at 1013 hPa. APFO starts to decompose above 105°C. At 20°C, the water solubility of PFOA is 9.5 g/l and of APFO > 500 g/l. When dissolved in water, the strong acid PFOA (pKa 2.5) dissociates. Pure PFOA at room temperature has moderate vapor pressure (2.3 Pa). The vapor pressure of APFO is much lower with 0.008 Pa. APFO or PF OA dissolved in water dissociate to ions. Although the dissociated fraction is not subject to volatilization, depending on the pH, pure PFOA might volatilize from wate r to a certain degree. Due to emissions for more than 50 years, PFOA is distributed worldwide in the marine environment, and hence may be transported to remote areas via the aqueous phase and the atmospheric phase. However, the significance of these sources are not currently kn own. Both atmospheric and aquatic transport mechanisms are actively being investigated. Possible substances subject to atmospheric long-range transport are PFOA precursors rather than PFOA itself. Potential precursors are fluorotelomer alcohols, -olefins and perfluoroalkyl sulfonyl derivates. These substances are degraded by OH radicals via gas-phase reactions to result partially in PFOA. The relative environmental significance of these sources is not known at this time. Due to the stability of the C-F bond, PFOA is pers istent in the environm ent. No degradation could be observed in th e studies on abiotic or bi ological degradability in water. Also the examinations on photolytic and photoche mical degradation in air indi cate high stability under environmentally relevant conditions. The half-life of the reaction with OH-radicals in the atmosphere is 130 days. According to the low adsorption potential and the water-solubility, PFOA is mobile in soil. The available ecotoxicological studies using APFO indicate a low acute t oxicity for aquatic organisms. In the short term tests using fish, invert ebrates, and algae, effective concentrations were as follows: Oncorhynchus mykiss LC 50 (96 h) = 707 mg/l (nominal) Daphnia magna EC 50 (48 h) = 480 mg/l (measured) Pseudokirchneriella subcapitata EC 50growth rate/biomass (72 h) > 400 mg/l (nominal)

6OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

The NOEC values determined in the chronic te sts using fish and daphniae, were 40 mg/l ( O. mykiss, NOEC, 85 d, measured) and 20 mg/l ( D. magna , NOEC, 21 d, measured). In a 10 day study using Chironomus tentans no effects were observed up to a nominal concentration of 100 mg/l. In addition to that, the following informat ion about effects on community level (indoor and outdoor microcosm studies) is available: Zooplankton community 35 d-LOEC species richness = 10 mg/l (nominal) Myriophyllum spicatum 35 d-EC 10 = 5.7 mg/l (measured) Myriophyllum spp. 35 d-NOEC = 23.9 mg/l (measured)
In several tests on effects using activated sludge , no inhibition of microbi al activity was measured up to a nominal concentration of 1000 mg/l. Concerning the effects on terrestr ial organisms, in a test using Caenorhabditis elegans without analytical verification, an EC 50 (48 h) of 973 mg/l (nominal) was calculated. In a chronic study using the same species, a reduction of abunda nce and egg production was observed at the 4 th generation at a concentration of 4.1x10 -3 mg/l (nominal). The NOEC for this endpoints was 4.1x10 -4 mg/l.

In tests with the rainbow trout Oncorhynchus mykiss a bioconcentration factor (BCF) of 0.038 and bioaccumulation fact ors (BAF) for organs of 27 (blood) , 8.0 (liver) and 4.0 (carcass) were obtained. These laboratory studies indicate a lo w bioaccumulation potential in fish. Some monitoring data suggest a low bioma gnification potential in aquatic food webs, while in some marine and Canadian Arctic mammalian food web studies a potential for biomagnification has been suggested. Further elucidation of the mechanisms leading to uptake and accumulation in biota is required.

Exposure APFO is used as a processing aid in the produ ction of fluoropolymers. In 2002, its world-wide production was about 200-300 metric tons. Entry in to the environment o ccurs during production and use of PFOA / APFO. Other sour ces for releases to the envir onment are residual contents of PFOA in fluoropolymer and fluoroelastomer pr oducts, PFOA as a byproduct in end products and fire-fighting foams containing perfluorocarboxyla tes, PFOA contaminations in perfluorooctyl sulfonyl (PFOS) based products, and PFOA contam ination in fluorotelomer products. An indirect source for PFOA in the environment is the de gradation (biotic and abio tic) of some fluorotelomer-based products. The global distribution of PFOA was demonstrated by several monitoring studies. Elevated PFOA concentrations were measured near indus trialized and urbanized regions. PFOA could be detected in air in concentrations in the range of pg/m3, ng/g dw in soil, in sediment, suspended matter, and sewage sludge. PFOA concentrations up to 67,000 ng/l and 3,200,000 ng/l were analysed in sewage effluent and landfill effluent. Sporadically, PFOA was determ ined in ground water samples (up to 3,400,000 ng/l). In fresh water samples (rivers, lakes, rain water) PFOA was regularily measured. The maximum concentration determined was 11,300 ng/l. Elevated concentrations of PFOA were also detected in coastal waters near industria lized and urbanized areas ; the maximum concentration was 15,300 ng/l.

In freshwater and salt water fish PFOA was detected occasionally. The maximum concentration (91 ng/g ww) was found in common shiner (liver sa mples) after a spill of fire retardant foam.

7OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

The highest PFOA concentration in birds was dete rmined in liver samples of cormorants (450 ng/g ww). However, it should be noted that for th is colony of cormorants the highest value (450 μg.kg -1 ww) appeared to be an outlier as the con centration was 4.5 times gr eater than the standard deviation of the mean. The occurrence of PFOA even in remote areas, was demonstrated by analysis of polar bear liver samples (highest concentration: 55.8 ng/g ww). Liver samples of other mammals (e.g. seals, whales, walrus, dol phin) contained PFOA; concentrations up to 62 ng/g ww.

In addition to the environmental measurements , PFOA was regularly anal ysed in human blood samples. While the pathways of human expos ure to PFOA and its sa lts are unknown, there are limited data on PFOA blood serum levels in both occupationally- and non-occupationallyexposed populations. It has beendetected in samples of humanblood (plasma, serum and whole blood), liver, seminal plasma, and breast milk fr om several countries throughout the world, including the US, Canada, Columbia, Poland, Belgiu m, India, Korea, Sri Lanka, Japan, Sweden and Germany. Preliminary US reports indicate that individuals living near a facility that uses PFOA have much higher PFOA serum concentrations than the levels previously reported for US populations. The routine finding of PFOA in human blood initiat ed research on the s ources of human exposure. Fluoropolymers and fluoroelastomers are typically manufactured through telomerization. As APFO is an essential processing aid in this process, trace amounts of PFOA may be generated as an unintended by-product. Fluorotelomer-based products ha ve numerous uses in many industrial and consumer products, including soil-, stain-, grease-, a nd water-resistant coatings on te xtiles and carpet, personal care products, and nonstick coatings on cookware, and uses in the automotive, mechanical, aerospace, chemical, electrical, medical, and building and construction industries. Consumer exposure to PFOA due to impurities in the finished products can not be excluded. However, based on analysis of consumer articles and using a simple co mpartmental model estimation, an explanation for the PFOA concentrations found in humans can not be given.

8OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

RECOMMENDATION AND RATIONALE FO R THE RECOMMENDATION AND NATURE OF FURTHER WORK RECOMMENDED

Human Health:
The chemical is a candidate for further work. The chemicals possess properties indicating a hazard for human health (eye irritation; subchroni c toxicity; potential carcinogenicity; developmental toxicity). In the US, data collected in the National Health and Nutrition Examination Survey (NHANES) will provide data on expos ure profiles of individuals across the U.S. Epidemiologic studies have not shown conclusively an associ ation of PFOA exposure and health outcomes but most of the studies were cross- sectional; further workis needed to understand any potential associations. Further work on the sp ecies differences in toxicokinetics and m ode of action to enhance our ability to predict risk in humans is currently underway. Member countries are invited to perform an exposure assessment and th en if indicated, a risk assessment.

Environment:
The chemical is a candidate for further work. PF OA is persistent in the environment. The primary environmental sink is the aqueous phase. PFOA tends to be dissipated into organisms and is eliminated from the body very slowly. Laborator y studies and monitoring data in some aquatic food webs indicate a low bioaccumulation potential in fish, but other data suggest a potential for biomagnification, e.g. in marine mammals and Canadian Arctic food webs. Hence, for substances like PFOA, bioconcentration values in fishmay not be the most relevant endpoint to consider. For PFOA, biomagnification may occur in air-breathing species (e.g., terrestrial mammals, birds and marine mammals). Further elucidation of the mechanisms leading to uptake and accumulation in biota is required.

The main industrial sources of environmental emi ssions appear to have been identified. Further research is needed to quantify the sources lead ing to the ubiquitous environmental distribution and human exposure. In conclusion, member countries are invited to pe rform an exposure assessment (consideration should be given to precurso rs to PFOA) and, if i ndicated, a risk assessment. Member countries are invited to consider risk management measur es, e.g. environmental emission reductions.

9OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID



## SIDS Initial Assessment Report

1.0 IDENTITY

1.1 Identification of the Substance

CAS Number: 335-67-1 3825-26-1

IUPAC Name: Pentadecafluorooctanoic Acid (PFOA) Ammonium Pentadecafluorooctanoate (APFO)

Molecular Formula: C 7 F 15 COOH C 7 F 15 COO NH 4 +

Structural Formula: COOH F FFFF F F

F FFFF F F F COOH•NH 3 F FFFF F F

F FFFF F F F

Molecular Weight: 414.07 g/mol 431.10 g/mol Synonyms: Perfluorooctanoic Acid; PFOA; Pentadecafluoro-1-octanoic acid; Perfluorocaprylic acid; Perfluoroheptanecarboxylic acid; Perfluoro-n-octanoic acid; Pentadecafluoro-n-octanoic acid; Pentadecafluorooctanoic acid; n-Perfluorooctanoic acid 1-Octanoic acid, 2,2,3,3,4,4,5,5,6,6, 7,7,8,8,8-pentadecafluoro Ammonium Perfluorooctanoate; APFO; C-8;Ammonium pentadecafluorooctanoate; Ammonium perfluorocaprylate; DS 101; FC 1015; FC 143; FX 1006; Fluorad ® FC 143; Perfluorooctanoic acid ammonium salt; Unidyne ® DS 101-20

There are a number of perflurorooc tanoate salts in commercial use but only in very small quantities (est. < 1 metric ton•yr-1). Examples are:

• Potassium perfluorooctanoate CAS # 2395-00-08

• Sodium perfluorooctanoate CAS # 335-95-5

• Silver perfluorooctanoate CAS # 335-93-3 These salts are considered in chapter “
2.2.1.3 Other Direct PFOA/APFO Sources ”. Other aspects of the salts different from APFO are not subject of this assessment report. For the purposes of this document, the anion of PFO (perfluorooctanoate) is frequently referenced as PFOA or APFO. APFO and PFOA ar e sometimes used interchangeably.

10OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

11 1.2 Purity/Impurities/Additives PFOA itself is mainly produced and used as amm onium perfluorooctanoate (APFO). The purity and impurity of APFO depends on the manufacturing pr ocess perfluorooctyl iodide oxidation and electro-chemical fluorination.

Industrial processes for the synthesis of PFOA
Electrochemical Fluorination (ECF) H(CH 2 ) 7 COF + HF + e - Æ F(CF 2 ) 7 COF Æ F(CF 2 ) 7 CO 2 NH 4 Octanoic acidfluoride APFO
Perfluorooctyl Iodide Oxidation F(CF 2 ) 8 I + [O] Æ F(CF 2 ) 7 COOH Æ F(CF 2 ) 7 CO 2 NH 4 Perfluorooctyl Iodide APFO

Process Perfluorooctyl Iodide Oxidati on Electrochemical Fluorination Purity of the commercial product: > 99% 1 99% 1

Impurities: • ≤ limit of quantification ( ≤ 0.01%) of each of the following acids PFNA, PFDA, PFUA, PFDDA, PFTDA • C-5 through C-7 isomers

• 0.2 % PFNA

• ≤ limit of quantification ( ≤ 0.01%) of each of the following acids PFDA, PFUA, PFDDA, PFTDA Linear / Branched 100% linear 2 < 30% branched (Simons, 1949; Kissa, 2001)

PFNA: perfluorononanoic acid, PFDA: perfluorodecanoic acid, PFUA: perfluoroundecanoic acid, PFDDA: perfluorododecanooic acid, PFTDA: perfluorotridecanoic acid

1 Puritiy after purification.

2 With regard to APFO used as a processing aid to manufacture certain fluoropolymers, the telomerisation process produces a range with different alkyl chain lengths. However the starting product perfluorooctyl iodide is refined and distilled to a high purity prior to production of AP FO. High purity APFO is n ecessary to produce PTFE.OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

1.3 Physico-Chemical properties

Table 1 Summary: Perfluorooctanoic Acid (PFOA) Physico-Chemical Properties

Property Value Reference IUCLID Substance type Organic compound 1.1.1 Physical state Solid 1.1.1 UV absorption in water no absorption < 290 nm Hori et al., 2005 Melting point (°C) 54.3 44 - 56.5 (6 references) Lide, 2003 Beilstein, 2005 2.1

Boiling point (°C) 188 (1013.25 hPa) 189 (981 hPa) Lide, 2003

Kauck and Diesslin, 1951 2.2

Density 1.792 g/cm 3 (20° C) HSDB, 2005 2.3 Vapour pressure (Pa) 4.2 (25° C) extrapolated from measured data 2.3 (20° C) extrapolated from measured data 128 (59.3° C) measured Kaiser et al., 2005; Washburn et al., 2005 Washburn et al., 2005

Washburn et al., 2005 2.4

Partition coefficient n-octanol/water (log value) *

Henry’s law constant ** Water solubility (g

⋅ L -1 ) 9.5 (25° C) Kauck and Diesslin, 1951 2.6.1 Solubility in organic solvents - pH value 2.6, 1 g/l (20° C) Merck, 2005 (reliability not assignable) 2.6.1

pKa 2.5

2.8
1.5-2.8 Gilliland, 1992, Ylinen et al., 1990 (reliability not assignable) Brace, 1962 Kissa, 2001 2.12

Critical Micelle Concentration (g

⋅ L -1 ) 3.6 - 3.7 Kissa, 1994

Conversion factor for the vapour phase 1 ppm = 17.21 mg/m 3

* For the surface active PFOA Kow is not measurable (USEPA, 2005). ** Henry’s law constant cannot be calculated from vapor pressure and solubility (dissociating substance).

12OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 2 Summary: Ammonium Perfluor ooctanoate (APFO) Physico-Chemical Properties

Property Value Reference IUCLID Substance type Organic compound 1.1.1 Physical state Solid 1.1.1 UV absorption in water no absorption < 290 nm (acid) Hori et al., 2005 1.1.2 Melting point (°C) 130 (decomposition) 157 - 165 (decomposition starts above 105° C) 3M, 1987 (reliability not assignable) Lines and Sutcliffe, 1984 2.1

Boiling point (°C) Decomposition Lines and Sutcliffe, 1984 2.2 Density 0.6-0.7 g/cm 3 (20° C) Grifith and Long, 1980 2.3 Vapour pressure (Pa) 0.0081 (20° C) calculated from measured data 3.7 (90.1° C) measured Washburn et al., 2005

Washburn et al., 2005 2.4

Partition coefficient n-octanol/water (log value) * 2.5

Henry’s law constant ** 3.3.2 Water solubility (g

⋅ L -1 ) at 20 °C > 500 Shinoda, Hato, and Hayashi ́, 1972 3M, 1987 (reliability not assignable) 2.6.1

Solubility in organic solvents (g

⋅ L -1 ) Heptane, Toluene : 0 Methanol, Acetone : >500 3M, 1987 (reliability not assignable) 2.6.1

pH value
in water at 23 °C approx. 5 3M, 1987 (reliability not assignable)
Begin of thermal decomposition (°C) 130

above 105 3M, 1987 (reliability not assignable) Lines and Sutcliffe, 1984
Critical Micelle Concentration ((g

⋅ L -1 ) (see table 1)

Conversion factor for the vapour phase 1 ppm = 17.92 mg/m 3

* Kow is not measurable for the surface active PFOA (USEPA, 2005). ** Henry’s law constant cannot be calculated from vapor pressure and solubility (dissociating substance).

13OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID



### 2.0 GENERAL INFORMATION ON EXPOSURE

2.1 Production Volumes and Use Pattern As stated before, PFOA is mainly producedas ammonium perfl uorooctanoate – APFO. The most important industrial processes for the synthesis of PFOA (as APFO) are Electrochemical Fluorination (ECF), and Perfluorooctyl Iodide Oxid ation. ECF starts with oc tanoic acidfluoride and hydrogen fluoride which are transforme d to APFO in multiple steps. In the Perfluorooctyl Iodide Oxid ation process, the perfluoroocty l iodide is oxidised to PFOA. PFOA is further transformed to its ammonium salt APFO which is the marketed form. Industrial processes for the synthesis of PFOA
Electrochemical Fluorination (ECF) H(CH 2 ) 7 COF + HF + e - Æ F(CF 2 ) 7 COF Æ F(CF 2 ) 7 CO 2 NH 4 Octanoic acidfluoride APFO
Perfluorooctyl Iodide Oxidation F(CF 2 ) 8 I + [O] Æ F(CF 2 ) 7 COOH Æ F(CF 2 ) 7 CO 2 NH 4 Perfluorooctyl Iodide APFO

2.1.1 Historical and Current Production
From 1947, the electrochemical fluorination (ECF) process was used worl dwide to manufacture the majority of APFO (Simons (1949)), and was 80-90% in 2000 (Prevedouros et al. (2006)). The largest production sites were in the U.S. (ECF pro cess ceased in 2002) and Belgium (production ceased in 2002), the next largest in Italy and small scale producers in Japan. From about 1975 to 2002, the remaining 10-20% of APFO was made by dire ct oxidation of perfluorooctyl iodide at one site in Germany (producti on ceased in 2004) and at le ast one site in Japan, and DuPont in the U.S (production started in 2002) . APFO has been mostly used in making fluoropolymers (see below). In 2002, DuPont and Daikin committed to selling APFO only for use in the fluoropolymer industry (SPI, 2002). Additional lim ited research quantity production, use and disposal have taken place in numerous academic and i ndustrial locations worldwide over the past fifty years as indicated by patents and papers in the scient ific literature (Pre vedouros et al., 2006; USEPA Dockets EPA-HQ-O PPT-2003-0012, EPA-HQ-OPPT- 2004-0112 and EPA-HQ-OPPT2004-0113). By 2002, the principal worl dwide APFO manufacturer by the EC F process, 3M, discontinued external sales and ceased production leaving only a number of relatively small producers in Europe and in Asia (OECD, 2004) compared to potential Ru ssian producers. New APFO production capacity based on perfluorooctyl iodide commenced in the U.S. by DuPont in late 2002 (DuPont, 2005a). At the time of shut down, 3M listed their APFO capac ity at 227 metric tonnes per year. Between 1992 and 2002, 3M’s total PFOA production averag ed 113 metric tonnes (250,000 pounds) annually (Wendling, 2003). Approximately 97% of the PF OA produced by 3M was used by its industrial customers and its own processes as a fluoropolymer processing aid, and the remainder was used in medical film coatings and electronics applicati ons. DuPont’s capacity was given as 136 metric ton14OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

nes per year at the time of st art up (USEPA docket AR-226-1094). Current producers are listed in Table 3. Estimated historical pr oduction is provided in Table 4.

Table 3 Current Manufacturers of APFO (DuPont 2005)

Region Producer(s) in 2005 Europe Miteni (Italy) some production in Russia
USA DuPont
South America None
Asia Daikin (Japan), several Chinese Producers

Table 4 Estimated Industry-Wide Global Historic APFO Manufacture

Time Period Number of Years Estimated Global Average Annual APFO Production (tonnes/year) Estim. Global Average APFO Prod. (cumulative tonnage over the period)
From 2003 3 Data not available Data not available
1995-2002 8 200 – 300 1600 – 2400
1980-1994 15 100 – 150 1500 – 2250
1965-1979 15 30 – 50 450 – 750
1951-1964 14 5 – 25 70 – 350
3600 – 5800* * Rounded to the nearest 100 t = metric tonnes data according to Prevedouros et al. (2006)

For information about emission reduction programes see Section “ 2.2.1.6 Measures and programes for the reduction of APFO/PFOA emissions ”.

2.1.1.1 Production in the European Union, Italy The only known producer of APFO in the European Union in 2005 is Miteni S.p.A. (owned by Mitsubishi Corp.). Miteni produces APFO at Trissino (Vincenza) using the electrochemical fluorination process (ECF). The actual production volume per year is in the range <100 tonnes. All of the volume produced is sold to customers outside the pr oduction site. An overview of Miteni’s historic PFOA production is listed in Error! Reference source not found. (Miteni, 2005).

15OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 5 Estimated Historic APFO Manufacture at Trisono, Italy

Time Period Number of Years Estimated Average Annual * APFO Production [t/yr]
2003-2004 2 40
1995-2002 8 30
1980-1994 15 20
1965-1979 15 10
1951-1964 14 0

* Rounded to the nearest 10 t metric tonnes (Miteni, 2005)

2.1.1.2 Production in the European Union, Germany Prior to 2004 virgin PFOA (as APFO) was produced in the Industrial Park Gendorf in Germany. The virgin APFO-production at the Gendorf site was ceased in 2004. The former production capacity was 5 to 40 metric tonnes/year. APFO was produced by oxidation of C 8 F 17 I. Virgin APFO production in former years is listed in Table 6.

Table 6 Historic APFO production at Gendorf, Germany (Dyneon, 2005)

2003 2002 2001 2000-1995 1994-1990
Production [tonnes/year] 5 10 40 20-25 5 – 15

Additional volumes of APFO were supplied by recovery and recycling from waste streams resulting in the reduced demand for virgin ma terial in recent years (Dyneon, 2005).

2.1.2 Current and Historical Use The largest direct use of PFOA is as the ammonium salt as a processing ai d in the manufacture of certain fluoropolymers. In this process APFO is used as an emulsifier in aqueous solution during the emulsion polymerisation of tetrafluoroethylene. APFO is not consumed during the polymerization process. The remaining APFO may be released into air or water, remain in the fluorpolymer product, be recycled, or be destroyed during preparation of the fluorpolymer (SPI, 2005). In 2002, there were thirty- two fluoropolymer manu facturing sites worldwide located in North America (eight), Japan (seven), China (seven), Europe (seven), Russia (two) and India (one) with a total fluoropolymer manufactur ing capacity of 144,000 metric t onnes (Prevedouros et al., 2006). The ammonium salt of PFOA, ammonium perfluorooc tanoate (APFO) is used in the production of most, but not all fluoropolymers. It is also possi ble to use other perfluor ocarboxylates as polymerization aids. For example for the production of Polyvinylidene fluoride (P VDF) some producers use ammonium perfluorononanoate (A PFN) instead of APFO. This depends on process and producer. The production and consumption of all fluoropolymers has grown st eadily and, along with it, the usage of APFO. Table 7 shows the worldwide c onsumption of fluoropolymers for 2001 (Ring, Kalin, and Kishi, 2002). An estimated 94,300 metric tonnes of fluoropolymers were consumed worldwide in 2001.

16OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 7 Global Fluoropolymer Production by Type for 2001, Excludes Chinese and Russian Production [thousand metric tonnes]

Fluoropolymer Grade US Western Europe Japan Totals

PTFE (polytetrafluoroethylene) 19 16 12 47FEP (fluorinated ethylene propylene) 10 1.5 3.7 15

PVDF (polyvinylidene fluoride) 11 6.1 1.2 19
Other 6.5 1.4 5.9 14
Total 47 25 23 95

Figures derived from Ring, Kalin, and Kishi, 2002.

APFO consumption can be estimated for the various fluoropolymers th at use it. The usage rate generally depends on the polymer type and the specifics of a given producer’s manufacturing process. Excluding PVDF, the consumption of APFO for othe r fluoropolymer types is estimated to be 200- 300 metric tonnes (Prevedouros et al., 2006). This range of estima ted consumption divided by the worldwide fluoropolymer production of 76,000 metric tonnes (94,300 metric tonnes less 18,300 metric tonnes PVDF consumption from Table 7) would yield a rough range of APFO usage factors of 0.27% to 0.40% on a dry polymer basis. Applying these factors to the es timated production volumes in Tabl e 7, a very rough usage pattern by main regions results. This is shown in Table 8 below for fluoropolymers excluding PVDF.

Table 8 Range of Estimated Global APFO Consumption for Fluoropolymers Production (excl. PVDF) for 2001 at Usage Factors Ranging from 0.27% to 0.40% (metric tonnes of APFO rounded to the nearest 5 tonnes)

Fluoropolymer Grade US Western Europe Japan Totals

PTFE (Polytetrafluoroethylene) 50 - 75 40 - 65 30 - 50 120 - 190FEP (fluorinated ethylene propylene) 25 - 40 5 - 10 10 - 15 40 - 65

PVDF (Polyvinylidene fluoride) N/A N/A N/A N/A
Other 15 - 25 5 15 - 25 35 – 55
Total 90 - 140 50 - 80 55 - 90 195 - 310

Note: Excludes Chinese and Russia Consumption

17OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Similarly, using historical production estimates and the usage factors ranging from 0.27% to 0.40%, the estimated historic usage fo r the manufacture of fluoropolymer s by major region, as shown in Table 9 below.

Table 9 Estimated Global Historical APFO Usage for Fluoropolymers Production since 1980 (excl. PVDF) at Usage Fa ctors ranging from 0.27% to 0.40% (metric tonnes of APFO rounde d to the nearest 50 tonnes)

Period US Western Europe Japan Totals
1980-1984 150 – 250 100 – 150 100 350 – 500
1985-1989 200 – 300 150 – 250 200 – 300 550 – 850
1990-1994 250 – 400 200 – 300 200 – 300 650 – 1000
1995-1999 300 – 500 300 – 400 300 – 400 900 – 1300
2000-2001 150 – 200 100 – 150 150 – 200 400 – 550
totals 1050 – 1650 850 – 1250 950 – 1300 2750 – 4200

Underlying fluoropolymer production volumes derived from Ring, Kalin, and Kishi (2002) (CEH Marketing Research Report).

Considering that these data ex clude Russia and China APFO cons umption, the total agrees well with the estimated range of APFO pr oduced for the same period in Table 4. The largest global manufacturers of fluoropolymers and fluoroelasto mers and their respective market share are listed in Table 10.

Table 10 Current manufacturers of fluoropo lymers and -elastomers, estimation of market share. (Not all manufacturers use APFO as a polymerisation processing aid.)

Company Rough estimation of market share in 2004 [%]

DuPont 27 Dyneon 16
Daikin 15
AsahiGlass 11 Solvay Solexis 10 DDE (JV Dow/DuPont) * 9
Atochem ** 4
Others 8 * DDE joint venture disbanded July 1st 2005 with fluoroelastomers remaining in DuPont at Dordrecht site (NL) ** Arkema: new name since 2005

18OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

2.1.2.1 Use as processing aid in the manufacture of fluoropolymers and fluoroelastomers in the EU Corresponding to the companies, as mentioned in Table 10, the manufacturing facilities for fluoropolymers and fluoroelastomers in the European Uni on are listed in Table 11. Downstream users of fluoropolymers (for example in the form of APFO containing dispersions) are not included.

Table 11 Fluoropolymers and -elast omers manufacturing sites in the EU (Plastics Europe 2005)

Company Manufacturing site, located in the EU

DuPont Dordrecht, Netherlands Dyneon Gendorf, Germany
Daikin Pierre-Benite (Lyon), France Oss, Netherlands Solvay Solexis Spinetta Marengeo, Italy Tavaux, France
Asahi Glass Thornton Cleveley s, Lancashire, United Kingdom
Arkema * Tavaux, France

* name before 2005: Atochem

2.1.2.2 PFOA for Non-Fluoropolymer Uses 3M had produced limited quanti ties of PFOA (approximately 3.6 metric tonnes (8,000 pounds) annually) for non-fluoropolymer use (Wendling 2003). The primary use of PFOA unrelated to the production of fluoropolymers occurred in medical f ilm applications where PFOA served as an antistatic additive in coatings sandwiched betwee n multi-layer medical films. In addition, a small amount (approximately 0.23 metric tonnes (500 pound per y ear), was derivatized as an ester to forma methacrylate co-polymer employed in the electronics industry as a humidity barrier in printed circuit boards and to secure silicon oil on precision bearings (Wendling 2003). PFOA was also used in a number of products as a surfactant i.e. cleaning fo rmulas, and dispersions were used in other minor uses with direct environmen tal releases i.e. lubricants.

2.2 Environmental Exposure and Fate

Direct and indirect sources of perfluorocarboxylates have been recently revi ewed (Prevedouros et al., 2006). A number of other studie s have considered the environmen tal fate of fluorotelomer alcohols as an indirect source of PFOA in the enviro nment (Dinglasan 2004, Ellis et al. 2004, Hurley et al. 2004, Lange et al. 2002, Martin et al. 2004c, Wallington et al. 2006, Wang et al 2005a, Wang et. al 2005b). Current understanding of the environmen tal behavior of PFOA and PFOA precursors is evolving and may change as new da ta are generated. These new data are also discussed in the following sections.

19OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

2.2.1 Sources of Environmental Exposure

2.2.1.1 Production of PFOA/APFO Current producers of the ammoni um salt of PFOA (called APFO) are listed in Table 3. There is minimal information available about emissions for producers in Europe and US A. No information is available for production outsi de the EU and USA. In 2005 the only production facility in the EU is the Miteni plant, located at Trissino, Italy. The former Hoechst and later Clariant pro duction at Gendorf, Germany, ceased in 2004. USA, Alabama, Decatur (3M) The 3M Corporation produced PFOA at various facilities in the US from 1969 to 2002 using electrochemical fluorination (ECF) technology. Current ly the 3M subsidiary Dyneon manufactures fluoropolymers at its Decatur, Alab ama plant and another company, Daikin America also manufactures fluoropolymers nearby. While not currently manufactured in Decatur, PFOA used in the manufacturing process of fluoropolymers is emitted to the surrounding air. It was estimated that approximately 5 kg (11 pounds) of PFOA per year are being emitted from the 3M Dyneon facility; the adjacent Daikin facility has air emissions of approximately 204 kg(450 pounds) per year (Wendling 2004 EPA-HQ-OPPT-2004-0112-0002). In addition, fr om 1978 to 1998, 3M incorporated sludge from the on-site wastewater treatment plant via subs urface injection in an on- site area designated as the Sludge Incorporation Area. It is not known how much total PFOA contaminated sludge was buried at this site; however, a USEPA hazard assessme nt for PFOA and its salts reported that less than 227 kg (500 lbs.) were buried in 1997 (USEPA 2002) . Treated process wast ewater containing PFOA was also discharged from th e facility to Baker’s Creek nearits confluence with the Tennessee River. The average PFOA concen tration in wastewater samples collected from the Decatur facility was 602, 766, 1,028, and 310 ppb in sampling data fr om 1998, 1999, 2000, and 2001, respectively (Santoro 2003). The average PFOA concentratio n in wastewater sample s at this location inJanuary and May 2003 were 58 and 88.3 ppb, respectively. For 1999, 3M estimated total PFOAcompound (PFOA and related salts) water releases of < 13.6 metric tonnes (30,000 lbs.) at its Decatur AL location (USEPA 2002). A data collecti on and monitoring program is underway and a screening level exposure assessment will characterize releases and exposures associated with this site. More data are being generated and will be made available to the public through the US EPA Docket in 2006 and 2007

. USA, Minnesota, Cottage Grove (3M) Analysis of PFOA in the wastewater effluent fr om the 3M Cottage Grove, MN facility indicated wastewater emissions of roughl y 20-2,000 ppb (Santoro 2003). The average concentration of PFOA in wastewater sample emissions from January -March 2000, September-October 2000, and December 2002 were 1,991, 216, and 180 ppb, respectively. The average PFOA concentration in wastewater samples from this facil ity for January-June 2003 ranged fr om 17.7 to 110.5 ppb (Santoro 2003). For 1999, 3M estimated total PFOA- compound wastewater releases of < 6.8 metric tonnes (15,000 lbs.) at its Cottage Grove MN location (USEPA 2002). For 1997, 3M estimated 0.885 metric tonnes (1,950 lbs.) of PFOA-compound stack releases at its Cottage Grove, MN location and another 2.0 metric tonnes (4,500 lbs.) from Cottage Grove incinerated offsite (USEPA 2002).

20OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

USA, North Carolina, Fayetteville (DuPont) Information about emission according to Table 12 is given.

Table 12 Emissions from DuPont APFO production sites (DuPont, 2005a) [metric tonnes /year]

APFO Emissions 2002 2003 2004
APFO Manufacture US <0.1 <0.1 <0.1 * Prior to 2002 DuPont did not manufacture APFO, it was manufactured by 3M (for 3M emissions data see above paragraph).

EU, Italy Trissino (Miteni) According to Miteni, information about production related emission of PFOA at Miteni’s production site Trissino to air or wate r is not available (Miteni, 2005). EU, Germany, Gendorf (Dyneon) Emission to Air / Water / Terrestrial Compartment: At the Gendorf site PFOA/APFO was produced in a closed process with marginal losses only. E.g. Emission to air before the shut down in 2004 was < 1kg/year. All emission quantities are included in the data provided in the use sec tion for the Gendorf site (Dyneon, 2005).

Other sites No information is available for the other producti on sites, for example in the UK, Poland, Russia, China, India or Japan). See Table 13 for the full list of fluoropolymer sites using PFOA as a polymerization aid.

2.2.1.2 Use of PFOA/APFO During the manufacture of fluoropol ymers, APFO is used to suspend and emulsify some fluoropolymers. APFO is not consumed but may be recycl ed or destroyed during preparation of the product, released into air or water or may remain in the product. One product containing a small amount of APFO is the aqueous fluoropolym er dispersion (AFD). AFD is used in the manufacture of products with the potential releas e of APFO during the manufacturi ng process (SPI, 2005). Based on 2003 survey data, approximately 15 % of the APFO used to make fluoropolymers world-wide remained in products sold to customers in the fo rm of AFD (SPI, 2005). Th e APFO content of commercially available AFD was reported to be typi cally <0.5% by weight (range of 0.7 to 0.2% by weight). In a survey undertaken by the Society of the Plastics Industry, In c. and the Fluoropolymer Manufacturers Group in the U.S.A., general inform ation on the manufacturing process and on-site sampling of representative proces sing facilities was collected to study APFO emissions from these sources. Several processes were studied including the manufacture of glass cloth coating and the manufacture/formulation of coati ng products for metal, glass, a nd specific polymeric substances. Common to these processes is a final step where th e polymer is heated in order to melt it onto the substrate, with the temperature dependent on the type of fluoropol ymer. However, steps until the final heating can release APFO through various waste streams. AFD is also used as an additive to dry cell batteries, in the manuf acture of valve and pump packing ma terials, in the manufacture and coating of film, manufacture of impregnated felt cloth for the manufacture of air pollution control

21OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

“bag houses”, and in the manufacture of fluoropolymer fibers to make a yarn, These manufacturing processes do not utilize a final high heat melting st ep. Results from this survey indicate that approximately 26% of the APFO present initially in th e AFD is released to air, water, and solid waste streams and 62% decomposes (durin g the heating steps). The remaining 12% APFO is used in low temperature processes and its fate was described as “undetermined” by the project (SPI, 2005). Main PTFE products sold in dry, bulk form (dry pow ders, granular resins) contain APFO in the order of 1-5 ppm (Cope, 1981).
Information about emission of APFO to the environment is only available for some sites. It is assumed that emissions of APFO take place at a ll fluoropolymer production sites. Some sites are named in Table 13. (In this tabl e only fluoropolymer production sites are listed not sites of fluoropolymer downstream users. At si tes of fluoropolymer downstream us er’s relevant APFO emissions also occur.)
The amounts of APFO used in 199 1-1992 are illustrated in the fo llowing Table 13 (USEPA, 1992- AR226-2423; USEPA, 1993 - AR226-2070).

Table 13 Global PFOA/APFO Users in 1991 and 1992 - historical

Site 1991 Amount [metric tonnes (lbs)] (USEPA, 1992- AR226-2423) 1992 Amount [metric tonnes (lbs)] (USEPA, 1993 - AR226-2070) DuPont, Washington Works, WV (USA) 32.3 (71,100) 30.5 (67,200)

Dordrecht Works (NL) 14.2 (31,400) 17.2 (37,950)
Shimizu 7.23 (15,950 6.67 (14,700) DuPont, Chambers Works, NJ (USA) 0.90 (2,000) 0.48 (1,050)

DuPont, Spruance, VA (USA) 0.68 (1,500) 0.64 (1,400) DuPont, Parlin NJ (USA) 1.13 (2,500) U.S. Major Customers 2.79 (6,150) Non-U.S. Major Customers 1.84 (4,060)

Table 14 Fluoropolymer manufacturing site using PFOA as a polymerisation aid in the manufacture of PTFE or FEP incl uding sites mentioned in the previous table (Will et al. 2005)

Country Name City

Americas
USA AGC Chemicals Americas Inc. (Asahi Glass) Bayonne, NJ
USA Solvay Solexis, Inc. Orange, TX
USA Daikin America Inc. Decatur, AL
USA DuPont Parkersburg, WV Europe
EU

22OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

23 Country Name City
Germany Dyneon 3M Gendorf Italy Ausimont S.p.A (now Solvay Solexis) Spinetta-Marengo
Netherlands DuPont Dordrecht

UK Asahi Glass UK (formerlyICI) Thornton Cleveleys
Poland Zak ł ady Azotowe w Tarnowie-Mo ś cicach S.A. Tarnow
Non-EU
Russia KCKK Kirovo-Chepetsk

Russia JSC "Halogen" Perm
Asia Japan Asahi Glass Ichihara Japan Daikin Industries Kashima Japan Daikin Industries Settsu, Osaka Japan DuPont-Mitsui Shimizu
China Chenguang Res Institute Zigong
China Fuxin Hengtong Fluorchemical Co., Ltd. Fuxin
China Jiangsu Meilan Chemical Group Corp. Taizhou
China Jinan 3F Fluorochemical Jinan
China Shanghai 3F Ne w Materials Shanghai
China Daikin Fluorochemicals (China) Co., Ltd. Changshu, Jiangsu
China Shangdong Dongyue Chemical Co., Ltd. Zibo, Shangdong
India Hindustan Fluorocarbons Ltd. Bashir Bagh ,Hyderabad

DuPont For 1999, DuPont estimated stack releases of 10.8 metric tonnes (24,000 lbs.) APFO at its Washington Works, WV location, plus another 7.26 metric tonnes (16,000 lbs.) from Washington Works that were incinerated offsite (USEPA 2002). For 1999, DuPont estimated the following APFO wastewater releases per locati on: Washington Works WV, 24.9 metr ic tonnes (55,000 lbs); Parlin NJ, 0.136 metric tonnes (300 lbs.); Spruance VA, 0.068 metric tonnes (150 lbs.); Chambers Works NJ, 4.3 metric tonnes (9,500 lbs.) (USEPA 2002). DuP ont also estimated that 1.8 metric tonnes (3,900 lbs.) of APFO in waste sludg e was landfilled on site in 1999 at their Chambers Works, NJ facility and 1.2 metric tonnes (2,600 lbs.) APFO wa s transferred offsite to a hazardous waste landfill from their Washington Works, WV plant. Aggregated emissions from DuPont producti on sites from 1999-2004 are listed in Table 15.OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 15 PFOA/APFO-emissions from DuPont sites (DuPont, 2005a) [metric tonnes /year]

APFO Emissions in metric tonnes 1999 2000 2001 2002 2003 2004

US 51 47.6 26.3 15.9 11.7 2
Asia 9.3 10.2 4.8 4.9 5.0 1.4 Fluoropolymer / Fluorotelomer Manufacture Europe (2 sites) 4.5 6.2 4.8 2.1 2.5 3.6

Total 64.8 64 35.9 22.9 19.2 7

Emissions from DuPont’s Dordrecht site located inThe Netherlands, Europe are listed in Table 16.

Table 16 APFO-emissions from DuPo nt Dordrecht site (DuPont, 2005a)

APFO [metric tonnes / year] 1999-2004
Air 0 – 2
Water 1 – 4
Product < 3Emission to terrestrial compartment (landfill) < 0.5

USA, West Virginia, Washington Works (DuPont) - historical A PFOA mass balance carried out for 1991 reported that 32.25 metric tonne s (71,100 lbs) of PFOA were purchased with an estimated disposal of 9.26 metric tonnes (20,410 lbs) and 5.10 metric tonnes (11,250 lbs) to the Ohio River and the Delawa re River, respectively, 2.08 metric tonnes (4,590 lbs) were emitted to air in the Ohio River valley, 1.7 metric tonnes (3,700 lbs) were released to land and 5.0 metric tonnes (11,000 lbs) and 0.02 metric tonnes (50 lbs) were disposed of in dispersions and fine powder, respectively. The release of PFOA to the Ohio Rive r can be separated into granular, fine powder/dispersion, and FEP sources at 0.072 (160), 1.3 (2,850), 7.9 metric tonnes/year (17,400 lbs/year), respectively. PFOA release to th e Delaware was entirely as a dispersion supernate. PFOA that was either recovered or destroye d in FP driers or in FEP ovens represented 3.7 metric tonnes (8,100 lbs) total ( 3.4 (7,500) and 0.32 metric tonnes (600 lbs), re spectively) (USEPA, 1992 AR226-2423). Emissions of PFOA from the Washington Works facility in 1983 were 9.84 (21,700), 5.76 (12,700), 3.5 (7,700) and 3.6 metric tonnes (8,000 lbs) to water, air, off-plant disposal, and to product, respectively. In 1986, these values were 8.66 (19,100), 7.35 (16,200), 3.8 (8,300), and 3.8 metric tonnes (8,300 lb s), respectively (Zipfel, 1987).
EU, The Netherlands, Dordrecht Work (DuPont) – historical An APFO mass balance carried out for 1990 reported that 14.2 metric tonne s (31,400 lbs) of PFOA were purchased with 9.62 (21,200), 3.22 (7,100), 0.18 (400), and 1.0 metric tonnes (2,300 lbs) disposed of to the Merwede River, to air, to the Do rdrecht municipal disposal area, and to product dispersions, respectively. 0.18 metric tonnes (400 lbs) (or 1.2 % of the feed) were either destroyed or recovered from FEP ovens (USEPA, 1992 AR226-2423).

24OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Japan, Shimizu Works (DuPont) – historical
A PFOA mass balance carried out for 1990 reported that of the initial purchase of 7.23 metric tonnes (15,950 lbs), 5.81 (12,800), 0.23 ( 500), 0.05 (10), and 0.77 metric tonnes (1,700 lbs) were released to Sugura bay, to air, to land, and to pr oduct with 0.43 metric tonnes (950 lbs) destroyed by incinerator (USEPA, 1992 AR226-2423).
USA, other historical emissions The DuPont Chambers Works facility reported that 0.71 (1,560), 0.16 (350), and 0.04 metric tonnes (90 lbs) of PFOA were released to the Delaware Ri ver, to air, and to product, respectively, in 1990. The Spruance facility reported th e release of 0.68 metric tonnes ( 1,500 lbs) of PFOA to the James River in 1990 (USEPA, 1992 AR226- 2423). A review of the PFOA em ission reduction to the environment reported that of 52.0 (114,600), 54.7 (120,650) , and 54.9 metric tonne s (121,100 lbs) used in 1990, 1991, and 1992, 87.4, 82.7, and 79.8% were releas ed to the environment, respectively (USEPA 1993 AA226-2070). 3M estimated that 0.885 metric tonnes (1950 lbs) PFOA compounds were released to the air from ve nt stacks from the Cottage Grove facility in 1997 with releases occurring from 100-200 days/year (ATSDR, 2005). EU, Germany, Gendorf works (Dyneon) About 15 % of the 2004 fluoropolymer industry wo rld market is produced by Dyneon. At the Gendorf site Dyneon’s throughput of AP FO at full production capacity utiliz ation is in the range of 10- 100 tonnes/year. A full production scale, advanced rec overy and recycling syst em is in operation at this facility. This has reduced emissions of APFO to the 1 – 10 tonnes/year range. A similar AFPO recovery system is in place for fluoropolymer disp ersion products. This ensures that they contain greatly reduced levels of APFO, thereby minimizi ng the releases of APFO at downstream processing locations. Manufacturers othe r than Dyneon also supply conve ntional fluoropolymer dispersions to downstream processors of these products (Dyneon, 2005).

Emission to air
Total releases of PFOA to air never exceeded 1 ton per year. Improvements to abatement and recovery technology have been routinely implemented to maintain a low air emission level even with increased production output. Emissions in 2005 were less than 0.5 tonnes per year.

• Before 1995 < 1 tonne/year

• 2005 <0.5 tonnes/year

Emission to Water Abatement and recovery technology has been continuously improved and implemented over the past few years.

• Before 1995 1-10 tonnes/year

• 2005 1-5 tonnes/year

25OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Emission with products (information for dispersions only) Conventional dispersion products contain some residual APFO, typically about 1000 – 2000 ppm. Recovery technology for removal of APFO to below 10 ppm was devel oped and installed by Dyneon in 2004. Many fluoropolymer dispersions are now being produced with this new technology. 2005 is a transition year. Therefor e an estimate of 2006 is provided.

• before 1995 1-10 tonnes/year

• 2005 <0.5 tonnes/year
Table 17 APFO emissions from Dyneon, Ge ndorf (Germany) production facility (Dyneon 2005)

APFO emissions in metric tonnes per year Before 1995 2005

Air < 1 < 0.5
Water 1-10 1-5
Product 1-10 < 0.5

2.2.1.3 Other Direct PFOA/APFO Sources

• Consumer and industrial products and fire-fi ghting foams containing perfluorocarboxylates including perfluorooctanoate (PFO): PFCAs and their derivatives have been used and are still present as additives and impurities in i ndustrial and consumer products as well as in fire-fighting foams (Kissa ( 2001), Moody et al. (2002) and Pr evedouros (2006)). As a consequence of residues and impurities of PFOA in some products and articles, use results in widespread release. Disposal of these products and articles at landfills is likely to lead to the release of PFOA to landfill effluents. Wastew ater treatment plant discharge can also be a source (Boulonger(2005).

• APFN: Ammonium Perfluorononanoate (APFN) is manufactured primarily in Japan by ozonolysis of a mixture of linear fluorotelomer olefins (Asahi Glass Co., 1975). The commerical product is a mixture comprised of mainlyodd acids 7-13 carbons in length. APFN is primarily used as a processing aid in fl uoropolymer manufacture, most notably polyvinylidene fluoride (PVDF).

• Silver (Ag) and potassium (K ) salts of PFOA have also b een items of commerce produced and used in very small quantities (approx. < 1 metric ton per year). 2.2.1.4 Indirect Sources (Chemical substances which contain PFOA as an impurity or may degrade to PFOA)
Chemical substances that include the C 8 F 17 moiety may be a source for PFOA. Examples are the telomer alcohols, sulfonates, sulfonamides, carbonates, phosphates, phosphinic and phosphonic acid (Dookhith, 2001; Kissa, 2001; Martin et al., 2004a; Be gley et al., 2005; Prev edouros et al., 2006). Manufacturing, use and degr adation of these derivates may be PFOA sources. Perfluoroalkyl precursor monomers Fluorotelomer-based products have been manufactured since the early-1970’s and used in many of the same industrial and consumer product applications as POSF-based products. Perfluoroalkyl iodide, Telomer A, is the basic raw material manufactured at one site in the U.S., one site in Germany and two sites in Japan. Residual fluorotelomer raw materials, such as 8:2 telomer alcohol, per26OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

fluoroalkyl iodide and perfluorosulfonamides are indirect sources of PFCA s. They may react during chemical processing or release to the e nvironment and form PFCA s (including PFOA). PFOA, known and suspected PFOA precursors are present in many fluorotelomer-based products. Environmental degradation of fluorotel omer raw materials present in fl uorotelomer-based products is a potential indirect source of PFC As. Microbial degradation, in vi vo metabolism in higher organisms and abiotic degradation have b een shown to produce low levels of PFOA (Wang, 2005a and 2005b; Fasano et al. 2006, Prevedouros et al., 2006). Recently, Dinglasan-Panlilio and Mabury (2006) identified the presence of residual unbound fluorotelomer alcohols in varying chain lengths (C6-C14) in seven commercially available and industrially applied polymeric and surfactant materials. It was determined that the fluorinated materials examined consist about of 0–1.6 % residual 8:2 Fluorotelomer alcohols of dry weight. Similarly, perfluoroalkyl sulfonamido based products, principally perfluorooctyl sulfonyl derivatives have been reported to contain residual raw materi al concentrations (e.g. sulfonamide alcohols, amides) at similar levels. (3M 1999c). Fluorotelomer based products Fluorotelomer-based products, monomers and poly mers, may contain trace levels of PFCAs, may contain PFOA (monomeric products up to % level) , may degrade to PFOA and may contain other impurities that may degrade to PFOA (e. g. Begleyet al. 2005, TRP, 2003, Washburn et al. 2005, Prevedouros et al., 2006). Impurities and unreacted monomers are present in telomer based polymer products. The 8:2 telomer alcohol is the most extensively studied of the unreacted raw materials pr esent in fluorotelomer based polymers. It can be transformed by both abiotic and biotic processes to form PFOA. Research suggests the existence of at least two bi odegradation pathways, (alpha and beta oxidation)by which, for example, the 8:2 telomer alcohol would biodegrade to either an even carbon numbered perfluorocarboxylic acid, i. e., perfluorooctanoic acid, or an odd carbon numbered perfluorocarboxylic acid, i.e., perfluorononanoic acid (Wang et al 2005a, Wang, et . al 2005b, Dinglasan 2004, Lange, et al. 2002). Wang et al (2005a) investigated the biodegradation of a 3- 14 C radiolabelled 8:2 telomer alcohol in a dilute domestic activated sludge test system fo r 28 days. Biodegradation products were separated and quantified by LC/ARC (on-line liquid chromat ography/accurate radioisotope counting). Transformation products were identified by quadrupole time of flight mass sp ectrometry. The alcohol strongly sorbed to the septa of the incubation containers with 41% of the mass of the added chemical found in septa extracts at the end of the study. However, the portion of the alcohol not sorbed to the septa was biodegraded to form PFOA and se veral intermediates. PFOA accounted for 2.1 percent of the total mass added. Other products included the 8:2 saturated acid at 27 percent of the total mass added and the 8:2 unsaturated acid 6 percent of the total mass added. Wang et al (2005a) suggest that this study indicates that in addition to the fo rmation of PFOA, alternative pathways of degradation exist for telomer alcohol s resulting in the mineralizatio n of the perfluorinated carbon bonds. This was confirmed in Wang et al. (2005b). In subsequent studies of the 3- 14 C radiolabelled 8:2 telomer alcohol , Wang et al (2005b), analyzed the headspace of sealed vessels containing mixed bacterial cultures and vessels cont aining activated sludge from a domestic sewage treatment plant unde r continuous air flow for up to 4 months. The mixed bacterial culture from an industrial wastew ater treatment sludge was enriched using 8:2 telomer alcohol. Using analytical methods previously described, a to tal of eight biodegradation products were identified, including three novel metabolites. No pe rfluorononanoic acid was observed in the test systems, suggesting that alpha-oxidation did not occur under the conditions of the test.

27OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

A similar observation was made by Wang et al. ( 2005a) and Dinglasan et al . (2004). In the mixed bacterial culture study, PFOA concentrations increased to day 56 and ultimately accounted for 6% of the total mass of the 8:2 te lomer alcohol after 90 days. 1- 14 C PFOA accounted for about 6% of the initial telomer alcohol after 56 days. In the continuous air fl ow and closed vessel activated sludge systems, 14 CO 2 was detected in the headspace, and increased up to 12 percent of the available 14 C over 135 days with periodic addition of ethanol . A net increase of fluoride ion in the mixed bacterial culture equivalent to 12 percent of tota l mineralization of the alcohol was also observed. Without additional ethanol, however, only 0.4% 14 CO 2 was reported during th e study period using an activated sludge from a different source. The auth ors concluded that with the regular addition of an organic carbon source, mineralization of multiple -CF 2 - groups was enhanced, and under conditions that may be present in a wastewater treatment plant, defluorination and mineralization of these groups to form shorter fluorinated carbon metabolites may occur.
When 14 C-PFOA was used as a starting material in a similar experiment, <0.6% of the 14 C mass balance was reported as 14 CO 2 after 28 days of incubation. This suggests that the significant quantities of 14 CO 2 reported in the 8:2 telomer alcohol experiment above using supplemental ethanol did not originate from the decarboxylation of 14 C-PFOA but, unlike from th e decarboxylation of telomer alcohol metabolites such as the unsaturated acid metabolites (Wang et al. 2005b). The presence of perfluorohexanoic acid in the mixed bacterial system, accounting for approximately 1 percent of the concentration of 3- 14 C radiolabelled 8:2 telomer alcoho l from day 28 to day 90 of thetest, also indicates that 8:2 telomer alcohol metabolites are undergoing defluorination and mineralization processes. Dinglasan et al. (2004) studied the transformation of 8:2 telomer alcohol in aerobic mixed microbial systems over an 81 day test period. The mixed cu lture was collected froma contaminated groundwater site and was initially enriched on 1,2-dichlo roethane. Degradation of the test substance was followed using gas chromatogra phy with electron capture GC/ECD. Transformation products were identified using either gas chromatography/mass spectroscopy GC/MS or liquid chromatography/tandem mass spectroscopy LC/MS/MS. By day 7, 85% of the 8:2 telomer alcohol (initially present at 50 μg/l) was degraded and by day 16, concen trations of the telomer alcohol were below the detection limit of 2 μg/l. By day 52, 6% of the to tal mass of 8:2 telomer al cohol was converted to PFOA; however by day 81 this value was reduced to approximately 3 percent of the total mass. Unlike the findings of Wang, et al . (2005b), the 8:2 unsaturated ac id was the most abundant metabolite comprising approximately 50 percent of the total mass of 8:2 telomer alcohol, with the 8:2 saturated acid accounting for less than 10 percent of the total mass of 8:2 telomer alcohol. In a study of Lange (2002) using a commercial mi xture of fluorotelomer alcohols containing primarily the 6:2, 8:2, or 10:2 forms, activated sludge from municipal wastewat er treatment plants was shown to nearly completely rem ove the 4:2, 6:2, 8:2, 10:2, and 12: 2 Fluorotelomer alcohols within 16 days, with moderate and litt le degradation for th e 14:2 and 16:2 alcohols, respectively. Disappearance was measured through analysis of the extract using HPLC/MS. No biodegradation occurred in the abiotic controls. Th e C5–C12 perfluorocarboxylic aciddegradation products were detected in the test cultures following the 16-day te st period. Only levels of perfluorooctanoic acid were quantified and were found at 6–7% of the to tal FTOH starting material or 23% of the 8:2 telomer. POSF-based products Perfluorooctyl sulfonyl-based (P OSF) products made by the ECF process contained PFCA impurities (3M, 2003a). POSF-based products have been ma nufactured at several sites in the U.S. as well at sites in Belgium, Italy, Germany and Japan. Pr oduction began in the 1950’ s. The principal (estimated 80-90% market share) manufacturer, 3M, phased out production starting in 2000 and completing its production phaseout in 2002. Global pr oduction by the major manufacturer of POSF as a

28OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

raw material in 2000 was approximately 3,700 tonnes. The global use, distribution and release of POSF-based products have been reported. Uses in cluded industrial applications and consumer products (i.e. food-contact paper, spray cans for do- it-yourself app lication) worldw ide. POSF-based products contained between 200-1,600 ppm of PF OA (solids basis). Although principally composed of fluorinated eight-carbon chains, the PFC A impurities in POSF-based product were a mixture of linear and branched (up to 30 wt%) chain isomers from fourto nine carbons in length. Degradation of POSF-based residual raw materials and products is also a potential indirect source of PFCAs in the environment. (3M 1977, 1978, 2003a, and Boulanger et al., 2005; Prevedouros et al. 2006). D’Eon et al. (2006) and Martin et al. (2006) estima ted that in the atmosphere perfluoroalkane sulfonamides may degrade to the re spective perfluorinated carboxylic acids in yields up to 45%. Recent studies simulating the atmospheric degradation of N-ethyl perfluorobutanesulfonamide (NEtFBSA), suggest that by analogy, N-ethyl perfluorooctanesulfonami de (N-EtFOSA) may also be a possible indirect source of pe rfluorocarboxylic acids, including PFOA in the environment (Martin et al. 2006). In smog chamber experiments usi ng N-EtFBSA, as an environmental analog for NEtFOSA, the atmospheric lifetime of N-EtFOSA was estimated as 20-50 da ys via reaction with photochemically generated hydroxyl radicals (Marti n et al. 2006). N-EtFBSA was employed in the experimental design of the study rather than N-Et FOSA, because it is muchmore volatile, thus facilitating its vapor-phase introduc tion into the smog chamber. The results of these experiments suggest that the degradation of perfluorobut anesulfonamides (and by analogy, N-EtFOSA) may serve as atmospheric sources of some perfluorocarboxylates, including PFOA. D’eon et al. (2006) conducted smog chamber experi ments in order to deduce the atmospheric chemistry of N-methyl perfluorobutan esulfonamidoethanol (N-MeFBSE). The results of this study indicated that the atmospheric degradation of N-MeFBSE is rapid and results inthe production of short chain perfluorocarboxylates among its degradation products. By analogy, it was concluded thatsimilar degradation mechanisms would occur for the perfluorooctane compounds Nmethylperfluorooctane sulfonamido ethanol (N-MeFOSE), which are wi dely distributed in the environment due to their previous use in oil and wate r repellent coatings for paper and packaging material. The authors of this study s uggest that N-MeFOSE is capable of contributing to PFOS and perfluorocarboxylic acid deposition that is observed in remote global locations. Biodegradation studies on N-EtFOSE ar e consistent with the suggestion that it is an indirect source of PFOA in the environment. In a screening test using activated sludge, N-Et FOSE was degraded to 34% of its initial concentration over 28 days (Rhoads et al., 2005) . The major degradation products were 2-(N-ethylperfluorooctane sulfonamido) acetic acid (N-EtFOS AA) and perfluorooctanesulfinate (PFOSI), which account for 60% and 15% of the metabolites , respectively. PFOS accounted for 7% of the metabolites at 28 days, N-EtFOSAaccounted for 4% of the metabolites and PFOA accounted for 2% of the metabolites . This work is in harmony with the results of Lange (2000). The biodegradability of N-EtFOSE (purity 99.9%) was measured in a 5-week screening study using wastewater treatment sludge as th e inoculum (Lange, 2000). At an initial concentr ation of 0.047 μg/ml, complete biodegradation was reported in 14 days while at an init ial concentration of 2.38 μg/ml, 90% biodegradation was reported in 35 da ys. The formations of metabolites over the study period were monitored using HPLC/MS. Ma jor metabolites included N-EtFOSAA and 2- (perfluorooctanesulfonamido) acetic acid accounting for 85% of the total metabolites by day 35.Minor metabolites included N-EtFOSA (0.1% of the products at day 35), perfluorooctanesulfonamide (FOSA, at 5.0%), perfluorosulfina te (3.5%), PFOS (7%), and PFOA (0.6%).

29OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Biodegradation of fluorotel omer-based polymers The potential of telomer based polymer products to form PFOA and other perfluorochemicals from degradation of the polymer is currently being ev aluated. A number of studies by industry, USEPA, and academics are underway examining the biodegra dation of fluorotelome r-based polymers with various Rf linkages in soil and sludge systems. Th ese studies will address questions about rates and mechanisms of degradation of fluorotelomer-based polymeric products. The potential for the degradation of the polymer linkage is currently being examined. Biodegradation of a fluorinated monomeric ac rylate (3,3,4,4,5,5,6,6,7,7,8,8,9,9,10,10-hept adecafluorodecyl methacrylate) has been reported by Dinglasan and Mabury (2005). This meth acrylate monomer was incubated with a sewage inocul um and fluorotelomer alcohol me tabolites measured using GC/MS and electrospray LC MS/MS. By day 48, the total measured telomer acids represented 3% of the initial mass of methacrylate monomer. Microbial - and fungal-mediated biodegradation of polyurethanes has been reported (How ard, 2002; Nakajima-Kambe et al., 1999). This process occurs quite slowly for unsubstituted polymers a nd the rate is expected to decreas e further with the presence of added substituents (Darby and Kaplan, 1968). A lthough data from long-term studies are not currently available for the fluorinated polymers, it is likely that the presence of perfluorinated substituents will increase the persistence of the polymer in the environment in comparison to an unsubstituted polymer by either physically blocking potentia l sites on the polymer b ackbone that might be available for microbial/fungal enzyme attack or th e perfluoro substituent will prevent polar enzymes from reaching the active enzyme site (Boese et al., 2005). In addition, the structure of the polymer backbone and the degree of subs titution will also determine its biodegradability. Polyester backbones suscep tible to esterase attack are expected to be considerably more degradable th an polyethylene structures (Nak ajima-Kambe et al., 1999). An inherent biodegradability screeni ng test (OECD 302B) was conducte d on a fluorotelomer-based acrylate polymer made predominantly from the fluorotel omer alcohol 2-perfluorooc tylethanol (Boese et al., 2005). More than 99.9% of th e 2-perfluorooctylethanol was bound as ester in the polymer with the remaining material reported as low molecula r weight residuals. Analysis of metabolites was conducted using GC/MS and LC MS/MS. Ester cleav age of the polymer was not reported after 28 days (Boese et al., 2005). The study was conducted using a municipal wastew ater treatment sludge inoculum under aerobic conditions over 28 days. Samples were taken from the system on days 0, 4, 14, and 28 and prepared using solid phase extraction foll owed by analysis using High Pe rformance Liquid Chromatography/Mass Spectrometry (HPLC/MS) for predicted biodegradation products based on earlier (proprietary) biodegradation studies. Due to the polymeric nature of the starting material the appearance of degradation products was measured rather than the direct measur ement of the loss of starting material. By day 4 of the study 114 % degradation of th e initial concentration of the starting material to a major biodegradation product was observed and mass balance appeared to have been achieved. Four minor metabolites were also observed. An incr ease in degradation products over 28 days in the abiotic controls suggests hydrolytic degradation also occurred. Biodegradation of the 8:2 fluorotel omer alcohol in soil to the 8: 2 saturated and unsaturated acids after 7 days has also been reported (Liu et al. 2005). Biodegradation of other polymeric structures 3M (2003d) has reported on the inherent aerobic aquatic biodegradation of a fluoroaliphatic polymeric ester. The specific chemical identities of th e test substance and metabolites were proprietary and not reported. In this study “fluoroaliphatic polymeric esters” were seen to be 100% degraded in a modified SCAS test within 4 days.

30OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

31 Studies of other polymer structures are bei ng conducted by US-EPA, Industry and academic researchers but to date but very limited preliminary data have been reported Recent papers by Sinclair and Kannan (2006) and others have reported that PFOA and other PFCAs are formed during wastewater treatment. Th e source of PFOA and other PFCAs is unknown, but their formation would be consiste nt with the degradation of polym er, 8:2 telomer alcohol, or other PFCA precursors. Indirect photolysis of precursors Fluorotelomer alcohols and pe rfluoroalkyl sulfonamido compounds such as perfluorooctyl sulfonamido ethanol, have been detected in the atmos phere (Stock et al. 2004, Shoeib et al. 2005). The ambient concentration of 8:2 telomer alcohol was reported to range approximately 9-123 pg/m 3 in North American air (Stock et al. 2004). Perfluoroalkyl sulfonamide concentrations were similar to telomere alcohols in North America. These substan ces have sufficient vapor pressure to be present in air at ambient temperatures (Kai ser et al. 2004, Krusic et al. 2005). The work of Ellis et al. (2004) and others (Martin et al. 2002, Hurley et al. 2004, Wallington et al. 2006) suggests that an additional indirect source of PFOA may be the at mospheric oxidation of fluorotelomer alcohols. Ellis et al. (2004) conducted smog chamber experime nts with mixtures of 4:2, 6:2, and 8:2 telomer alcohols using Cl atoms in the ab sence of NOx to initiate the oxi dation of fluorotelomer alcohols. Cl atoms are believed to share a common mechanis m with OH radicals for attack on fluorotelomer alcohols (the hydrogen atom abstraction from th e -CH2- group bearing the alcohol functionality) and are believed by the authors to be a reasonable surrogate for OH radicals in atmospheric oxidation studies of Fluorotelomer al cohols. Cl atoms were employed in the smog chamber experiments because reaction of telomer alcohols with Cl pro ceeds at a much higher rate as compared to hydroxyl radicals thereby facilitati ng the ability to analyze the degr adation products more readily. The results of the experiments demonstrated that the fluorotelomer alcohols we re oxidized to form perfluorocarboxylic acids. The overall yield of perfluorocarboxylic acid s at an intermediate time point in the degradation was approximately 5 percent with PFOA produced at carbon yields of approximately 1.5 percent. The authors noted that in th e presence of NOx, as in polluted urban air, fluorotelomer alcohols will undergo a different series of reactions th at do not directly lead to perfluorocarboxylic acid formation. However, in remote areas where the concentrations of HO 2 and other peroxy radicals are comparable to those of NO and NO 2 it is likely that reactions leading to perfluorocarboxylic acid formati on will play an important role. The atmospheric lifetimes calculated by Ellis et al . (2004), Martin et al. (2002), and Hurley et al. (2004b) using smog-chamber studies on the gas-phase atmospheric fate indica te that atmospheric long-range transport of the fluorotel omer alcohols is possible. In addition, the studies demonstrate that these precursor substances can form PFCAs as a result of atmospheric chemical degradation processes. For instance, gas-phase hydroxyl radical degradation of fluorotelomer alcohols yields 1- 10% PFCA (Ellis et al. 2004) while the perfluoroalkyl sulfonamido substances gave PFCA yields up to 45% 3 (Martin et al. 2006, D’eon et al. 2006). Th e authors proposed that gas-phase atmospheric degradation of precursor substances in remote areas may be a source of PFOA and other PFCAs in polar and remote regions far from di rect PFOA sources. (Ellis et al. 2004). 3-D global atmospheric transport modeling computing showed an estimated annual PFOA flux to the arctic of 0.4 tonnes per year (Wallington et al. 2006). In an unpublished presentation on global steady-state

3 The authors of Martin et al. (2006) state regarding Ellis et al. (2004): "...for fluorotelomer alcohols ... 14% of the perfluorocarbons are ultimately expected to be incorporated into PFCAs ..."OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

multi-media fate modeling work, other researchers have suggested that 8:2 fluorotelomer alcohol preferentially partitions to air and has the pote ntial for long range transport to Arctic regions (Klasmeier et al. 2004). An atmospheric degradation mechanism for 8:2 te lomer alcohol has been proposed and is summarized in Figure 1 (Wallington et al. 2006).

Figure 1 Proposed simplified atmospheric degradation pathway for 8:2 telomer alcohol (Wallington et al. 2006, modified).
Oxidation of 8:2 telomer alcohol proceeds by react ion with hydroxyl radicals and leads to the formation of perfluoroaldehyde in 100% yield, which also reacts w ith hydroxyl radicals and atmospheric oxygen producing acyl peroxy radicals (C 8 F 17 CH 2 C(O)O 2 ). It is assumed that these radicals react with NO, NO 2 and hydroperoxy radicals lead ing to the formation of various end products. The authors reported these modeling results sugges t the degradation of fluorotelomer alcohols contribute to PFCAs observed in remote locations. Th ey concluded that it is likely that PFNA, PFOA, and other shorter PFCAs are significant products from 8: 2 fluorotelomer alcohol oxidation and that

32OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

this can be a significant global source of perfluorocarboxy lic acids. As a result of this oxidation, PFOA is likely to be ubiquitous in the NorthernHemisphere. The yield of PFOA and other PFCAs are of the correct order of magn itude to explain observed PFCAs levels in Arctic fauna (Wallington et al. 2006). The gas-phase chemical degradation pathway for fluorotelomer alcohol ha s been well defined and investigated (Figure 1). However, additional res earch indicates that additional atmospheric processes may play an important role in the overall atmospheric fate of fluorotelomer alcohols. Fluorotelomer alcohols, while inde ed volatile, may be removed to some extent from the gas-phase due to their strong adsorption properties. In partitioning studies, poly parameter linear free energy relationships were derived for fl uorotelomer alcohols and olefins. The study concluded that environmental partitioning properties of fluorotelomer alcohols and olefins cannotbe accurately predicted from correlations with exis ting values of Kow or vapor pre ssure. Interface sorption should be taken into account and modeling partitioning to bulk compartments does not accurate ly predict behavior. Partitioning to air may be 1-2 orders of magnitude highe r then currently measured data suggest. (Goss et al. 2006). Heterogeneous uptak e studies found that gas- phase 8:2 fluorotelomer alcohol was rapidly taken up by oc tanol droplets suggesting that partitioning to organic-containing cloud/fog droplets and aerosol particles may be an atmospheric loss mechanism (Li et al. 2006). Recent smog-chamber studies on perfluoroaldehyde s, a key transformation product in the atmospheric pathway shown in Figure 1, s how that perfluoroaldehyde photolysis may be an atmospheric degradation pathway (Chi appero et al. 2006).

Telomer alcohols have been detected in air in No rth America (Shoeib et al ., 2006) and in precipitation (Scott et al. 2006b, DuPont, 2005). Existing data is insufficient to understand the transport and removal processes which control the behavior of these chemicals in the atmosphere. In a different study, unreacted 8:2 telomer alcohol has also been shown to undergo indirect photolysis in natural waters yielding PFOA (Gauthier and Mabury 2005). Sim ilar to the degradation pathway in the atmosphere, the degradation of 8:2 telomer alcohol and the subsequent production of PFOA is facilitated by hydroxyl radicals present in the water. In distilled water containing 10 mM hydrogen peroxide exposed to artifi cial sunlight, the photolysis half -life of 8:2 telomer alcohol was 0.83 ± 0.20 hours, and PFOA was formed at approximate ly 35% of the initial amount of 8:2 telomer alcohol (Gauthier and Mabury 2005) . Using a lower initial starti ng concentration of 0.1 mM hydrogen peroxide resulted in a longe r half-life (38.2 ± 6 hours), and th e PFOA yield was about 12%. In similar experiments using natural water obtained fr om Lake Ontario, the half-life of 8:2 telomer alcohol was 93.2 ± 10.0 hours and the PFOA yield was approximately 7%. In recently published studies, th e gas-phase atmospheric chemistr y of perfluoroalkyl sulfonamido substances was investigated and PFCAs were detected as products of the atmospheric chemistry. Perfluoroalkyl sulfonamido substanc es are thus a plausible source of PFCAs (Martin et al. 2006, D’eon et al. 2006). The qua ntitative contribution to PFOA presen ce in the environment and in remote regions has not been determined. Thermal degradation of fluorinated compounds Thermal degradation of fluoropolymers may result in formation of monomers and other PFOA precursors (Ellis et al. 2001). Perf luorinated carboxylic acids were identified as minor combustion products following the thermal degradation of pol ytetrafluoroethylene (Ellis et al. 2001). Some fluoropolymers lose greater then 1% mass/hour in air at maximu m processing temperatures. The principal evolved gas in one study was carbonyl fluoride (COF 2 ) (Baker and Kasprzak, 1993). In a study of Leung et al. (2002) thermal degradat ion of FEP (fluorinated ethylene propylene) showed to produce 50 ppm PFOA and the C6 perfl uorocarboxylic acid each. For PTFE there were no findings (LOD not noted).

33OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Other studies indicate that no PFOA was formed under simulated typical municipal incineration conditions of fabrics treated with fluorotelomer-b ased polymers (Yamada et al. 2005). Additional research is ongoing to help clar ify the behavior in incineration (USEPA 2006). On April 16, 2003 an Enforceable Consent Agreement (ECA) was en tered into by USEPA a nd several fluoropolymer manufacturers (AGC Chemicals Am ericas, Inc., Daikin America, Inc., Dyneon, LLC, and E.I. du Pont de Nemours and Company). Th e purpose of this ECA was to de velop a laboratory-scale incineration testing program of fluoropolymers (USEPA 2003b USEPA Docket OPPT-2003-0071). The fluoropolymers to be tested in clude: Dry non-melt fluoropolymer resin (CAS 9002-84-0); Dry nonmelt fluoroelastomer gum (CAS 27029-05-6) ; Aqueous dispersion (CAS 9002-84-0). Begley et al. (2005) determined the PFOA content of PTFE film of non-stick cookware before and after heating empty pans to 320 °C. The PFOA cont ent decreased from 4-75 μg/kg to below LOD. The authors conclude that significant amounts of PFOA are not generated and remain in the cookware after an extreme heating event. 2.2.1.5 Historical global estimation of sources One estimation of the historical global environmental releases of PFCAs fr om indirect sources and direct sources has recently been compiled. These data and the correspondi ng sources are presented in Table 18 (based on Prevedouros et al., (2006). The snapshots of PFCA environmental releases in 2000 by carbon chain length shown in Figure 2 and Figure 3 were taken from this paper. Table 18 Estimated Historical Global Releases of Perfluorocarboxylic Acids (PFCAs) based on (Prevedouros et al. 2006)

Environmental Input Source Historical Time Period (years) Estimated Total Global Historical PFCA Emissions (tonnes)
Other Direct PFCA Sources Fluoropolymer Manufacture (APFN) 1975-2004 400 - 1,400 Fluoropolymer Processing (APFN) 1975 - 2004 10 – 20 Aqueous Fire Fighting Foams (AFFF) 1965-1974 50 – 100

Consumer and Industrial Uses 1960-2000 40 – 200

Indirect PFCA Sources

POSF-Based Products 1960-2002 PFCA Residual Impurities 1960-2002 20 - 130 POSF-Based Precursor Degradation 1960-2002 1 - 30

POSF-Based AFFF 1970-2002 3 - 30

Fluorotelomer-Based Products 1974-2004 PFCA Residual Impurities 1974-2004 0.3 - 30

Fluorotelomer-Based Precursor Degradation 1974-2004 6 - 130

Fluorotelomer-Based AFFF 1975-2004 < 1

34OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

57% Fluoropolymer Manufacture - APFO 17% Fluoropolymer Manufacture - APFN
10% APFO Manufacture 5% Fluoropolymer Dispersion - APFO 4% PFCA AFFF, Industrial & Consumer Uses
3% APFN Manufacture 1% POSF Products - PFCA Impurities 1% Fluorotelomer Raw Material Degradation 0.3% Fluorotelomer Products - PFCA Impurities 0.3% Fluoropolymer Processing - APFN 0.3% POSF Raw Material Degradation
0.3% POSF-Based AFFF 10%

5%

4%

3%

0.3% 0.3% 1%

0.3% 0.3%

1% Fluoropolymer Manufacture - APFO 57 % 17 % Fluoropolymer Manufacture

- APFN 10 % APFO Manufacture

Figure 2 Global Historic PFCA Emission s by Source (Prevedouros et al., 2006)

#Carbon Acid Metric Tonnes 6 1 8 200 9 25 10 2 11 7 12 0.3 13 2 

### 232 135 10

6 13 12 9 8

11

Figure 3 Estimated 2000 Global PFCA Emission s by Chain Length (Prevedouros et al., 2006)

2.2.1.6 Measures and programes for the re duction of APFO/PFOA emissions In March of 2006, the US Environmental Protection Agency announced the EPA 2010/2015 PFOA stewardship program. The particip ation requires volunta ry global corporate commitment to two goals: 1) A commitment to achieve, no later than 201 0, a 95% reduction - measured from a year 2000 baseline - in both facility emissi ons to all environmental media of PFOA, precursor chemicals that can break down to PFOA, and related higher homol ogue chemicals. In para llel, it requires an

35OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

equivalent reduction in product content levels of PFOA, precursor chemicals that can break down to PFOA, and related higher homologue chemicals. 2) A commitment to working toward the elimination of PFOA, PFOA precursors, and related higher homologue chemicals from emissions and products by five years thereafter, or no later than 2015. A number of projects are proposed or under way worldwide to achie ve emissions reductions from fluoropolymer and fluorot elomer manufacturing. Fluoropolymer manufacturers have recently installed additional capa bility to capture and recycle APFO. The global APFO manufactur ing emissions have decreased from about 45 tonnes in 1999 to about 15 tonnes in 2004 and are expect ed to be 7 tonnes in 2006 (FMG, 2002). Aqueous APFO solutions have been processed for reuse in al l regions. Dramatic (>90%) reductions in overall APFO/PFOA environmental releases over the past few years have been reported (DuPont, 2005a; Dyneon, 2004). The recent historical trend in APFO emissions from fluoropolymer manufacture shows a decrease from hundreds of tonnes in 1999 to le ss than 50 tonnes proj ected in 2006 (Prevedouros et al., 2006). For example, DuPont’s Wash ington Works PFOA rec overy program has resulted in a recovery of the 20% solution of C- 8 from 5.0 metric tonnes (11,000 lbs) in 1992 to 58.0 metric tonnes (130,000 lbs) in 1997 (DuPont, 1998 – AR226-2084).

2.2.2 Photodegradation Direct photolysis of APFO was examined in two separate studies (3M, 1979; 3M, 2001b) and photodegradation was not observed ineither study. In the 3M (1979) study, a solution of 50 mg/l APFO in 2.8 liters of distilled water was exposed to simulated sunlight at 22±2 oC. Spectral energy was characterized from 290-600 nm with a max output at ~360 nm. Direct phot olysis 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. This was also done with older, much less sensitive analytical methods. Current methods may be able to detect degradation products. In the 3M (2001b) study, both direct and indirect photolysis were examined utilizing techniques based on USEPA and OECD guidance documents. To determine the potential for direct photolysis, APFO was dissolved in pH 7 buffe red water and exposed to simulate d sunlight (Scrano et al., 1999; Nubbe et al., 1995). For indirect photolysis, APFO was dissolved in3 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 phot odegradation of APFO. One matrixwas a pH 7 buffered aqueous solution containing H 2 O 2 as a well-characterized source of OH radicals (Ogata et al., 1983; Lunak and Sediak, 1992). This tested the propensity of A PFO to undergo indirect photolysis. The second matrix contained Fe 2 O 3 in water that has been shown to generate hydroxyl radicals via a Fentontype reaction in the presence of natural and artificial sunlight (Kachanova and Koslov, 1973; Behar et al., 1966). The third matrix contained a standard solution of humic materi al. 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 quantitat ion. There was no conc lusive evidence of direct or indirect photolysis whose rates of de gradation are highly depende nt on the experimental conditions. Using the iron oxide (Fe 2 O 3 ) photoinitiator matrix model, the APFO half-life was estimated to be greater than 349 days. According to Hori et al. (2004a, 2005) aqueous solu tions of PFOA absorb light strongly from the deep UV-region to 220 nm. A weak, broad absorpti on band reaches from 220 to 270 nm (no absorption coefficient stated). The irra diation of a 1.35 mM PFOA solution (29.6 μmol) in water with light from a xenon-mercury lamp (no radiant flux stated ) for 24 hours resulted in a ca. 40 % reduction (12.8 μmol) of PFOA concentra tion. Concentrations of CO 2 and fluoride increased simultaneously.

36OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Small amounts (0.1-5 μmol) of short ch ain perfluorinated hydrocarbon acids (C 2 -C 7 ) were detected. After irradiation about 12 % of total carbon and fluoride couldn’t be quantified. Addition of highpressure oxygen (4,800 hPa) and hydrogen peroxide increased the amount of fluoride and carbon dioxide detected. The addition of the phot ocatalyst tungsten heteropolyacid ([PW 12 O 40 ] - ) or persulfate (S 2 O 8 2- ) (Hori et al., 2005) accelerates the reaction ra te. Due to the short wave length used for irradiation (< 300 nm) the photodegr adation described may be of lim ited relevance for an aqueous environment but may be used as a technical process. Hori et al. (2004a) suggest using the degradation for the treatment of PFOA contaminated industrial wastewater. The photodegradation is not considered a pathway for environmental degradation of PFOA (absorption of light with wavelength below 300 nm). Hurley et al. (2004) determined the rate constant s of the reactions of OH radicals with a homologous series of perfluorinated ac ids (from trifluoroace tic acid to nonafluoropentanoic acid) in 700 Torr of air at 296 K. For C 3 to C 5 chain length had no discernible im pact on the reactivity of the molecule. The rate constant k(OH + F(CF 2 ) n COOH) = (1.69±0.22) × 10 -13 cm 3 molecule -1 s -1 for n = 2, 3, 4, respectively. Atmospheric lifetimes of F(CF 2 ) n COOH with respect to reaction with OH radicals are estimated to be approximately 230 days for n = 1 and 130 days for n > 1. (Calculation of lifetime by comparison with CH 3 CCl3 (half-life 5.99 years, k = 1.0×10 -14 cm 3 molecule -1 s -1 ).) The authors conclude, that the major atmospheric loss mechanism of perfluorinat ed carboxylic acids is dry and wet (particle mediated) de position which occur on a time scal e which is probably of the order of 10 days. Reaction with OH is a minor atmo spheric loss mechanism for perfluorinated carboxylic acids.

Conclusion
PFOA has not been shown to under go direct photodegradat ion at environmenta lly relevant conditions (wavelength of light > 300 nm ). There is an indication (conclu sion by analogy) of slow indirect photodegradation in air (half-life 130 days).

2.2.3 Stability in Water The 3M Environmental Laboratory (2001a) performed a study of the hydrolysis of PFOA. The study procedures were based on USEPA’s OPPTS Guideline Document 835.2110 (USEPA, 1998); although the procedures do not fulf ill all the requirement s of the guideline, th ey 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). During the study, samples were prepared and examin ed at six different pH levels from 1.5 to 11.0 over a period of 109 days. Experiments were perfor med 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 da ta quality objective and were rejected. Also rejected were the data obtained for pH 1.5 becau se ion pairing le d 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 depend ence 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 mo st 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.

Conclusion PFOA is hydrolyticallystable in water.

37OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

2.2.4 Transport between Environmental Compartments

2.2.4.1 Model calculations An octanol-water partition coefficient (Kow ) cannot be determined for the surface active PFOA. Therefore calculating the distri bution between environm ental compartments with models using Kow (for example the Mackay model) is not applicable.

2.2.4.2 Volatility Different studies have reported di fferent values for PFOA vapor pr essure. The vapor pressure of PFOA was determined experimentally from 60° to 190° C, using a dynamic procedure where the equilibrium temperature is measured at controlle d pressure (Kaiser et al . 2005). The extrapolation of the measured pressure to lower temperature was based on a fitted Antoine equation. For 20° C a vapor pressure of 2.3 Pa was extrapolated.The vapor pressure of APFO was determined based on a standard procedure described by the USE PA in which the quantity of APFO transported by a known volume of a carrier gas is measured. Measur ements reached from 90° to 165° C. A vapor pressure of 0.008 Pa was extrapolated for 20°C (W ashburn et al. 2006) with a fitted Antoine equation. Studies with wash bottles (impinger studies) were performed to examine the volatility of aqueous APFO solutions (3M, 1993). Solutions of APFO containing ammoni um acetate in water/1-propanol (50:50) or phase transfer agents , e.g., n-alkyldimethylbenzylammoni um chloride were blown with 280 liters of air at a flow rate of 1 L/min. (3M, 1993). The results indicate there is some loss of APFO, but most of the solutions retained over 80% or more of the APFO. The average retention was 92% for APFO, indicating that some APFO disappeared out of th e solutions. However, some of the solutions, particularly the n-alkyldimethylbenzylammonium chlo ride solution, appear to retain all the fluorochemicals. (3M, 1993). The authors concl uded that it is unlikelythat this fluorochemical was removed by bubbling air through water due to its very low vapor pressure. Authors suggested that a more plausible mechanism for loss fro m the solution phase is concentration of the surfactants in foam and loss from the bubbled solutions as foam or micro-droplets. The physical-chemical properties of PFOA and its an ion PFO are distinctly different and govern the expected environmental fate and tr ansport. PFOA is an acid with a pKa of about 3 (s ee chapter “1.3 Physico-Chemical properties”). At pH values of 4.5, approximately 97% of this chemical is present as the perfluorocarboxyla te anion. The remaining 3% are prot onated. In the aqueous phase, at pH values between 6 and 8, lower amounts of PFOA (< 3%) are in equilibrium with PFO. However, the protonated form has sufficient vo latility to leave surface and atmo spheric water and/or soil, and generating a slow release of PFOA into the atmos phere. The environmental relevance of this release is unknown.

Conclusion While perfluorooctanoate, the anionic form, is no t volatile, pure PFOA (prot onated) is moderately volatile. When dissolved in water the strong acid PFOA dissociates. The degree is dependent on the pH. Environmental conditions control par titioning between environmental media

2.2.4.3 Long-Range Transport PFOA, as the anion perfluorooctanoate, PFO, has been detected in remote areas of the world in monitoring programs involving various abiotic and biotic samples. For example, PFOA has been

38OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

measured in air (Shoeib et al. 2006) and biota such as polar bears and seals in the Canadian Arctic (see section 2.2.8b Environmental Monitoring, DeSilv a and Mabury 2004, Martin et al. 2004c). No information is available about curre nt or historical use of PFOA or related subs tances in the Arctic. A possible explanation for this fi nding is the long-range transport of either PFOA orpotential precursors. Two possible transportation pathways include atmospheric and aquatic transport. Atmospheric Transport Due to the relative vapor pressures of APFO, PFOA, and PFO, the chemical form potentially most subject to gas-phase atmospheric transport is PFOA. Franklin (2002) suggested that in the presence of water in air (humidity), gase ous PFOA condenses to aerosol part icles and dissociates to the corresponding perfluorooctanoate, resulting in a low vapour pressure. According to Franklin (2002), the atmospheric lifetime of PFOA (respectively its salts) was calculated in the order of days when emitted from a ground source.
Additional sources of PFOA to the atmosphere are the degradation or transformation of precursors, which could lead to indirect environmental releas es. Potential precursors in clude related fluorinated chemicals which are detectable in the atmosphere (e.g., fluorotelomer alcohols, olefins, and perfluoroalkyl sulfonamido substances) which can degrade in the atmosphere or after deposition to the surface to PFOA (precursor substances and degradation mechanisms see section “2.2.1.4 Indirect sources”). Calculations using a three-dimensi onal global atmospheric chemistry model (IMPACT) indicate that 8:2 fluorotelomer alcohol (widely used in industrial and consumer products) degrades in the atmosphere to give PFOA (Wallington et al ., 2006).. FTOHs have suffici ent vapor pressure to be present in air (Prevedouros et al., 2006). Smog chamber studies prove the potential for FTOHs to react in the atmosphere with ubiquitous OH radicals to yield PF OA (Ellis et al., 2004). Telomer A ( ≥ C8) is the initial produc t for the manufacturing of C8 telomer alcohols. El lis et al. (2004) showed that the atmospheric lifetime of short chain FTOHs, as determinedby their reaction with OH radicals, was approximately 20 days. However, there is not en ough data available to estimate how muchthe different sources contribute to the PFOA detected in the Arctic and in biota ofremote areas. While there is evidence for the possible role of precursors for the l ong-range atmospheric tr ansport of PFOA, the extent to which these precursors and their transformation may explain th e concentrations of PF OA found in remote areas such as the Canadian Arctic is uncertain. Aquatic Transport and Marine Aerosols Another possible mechanism for the transport of PF OA to the Canadian Arctic is aquatic transport (Prevedouros et al., 2006). Given PFOA’s environmental persistence, high water-solubility and the fact that PFOA and related substances have been emitted to air and water for approximately 50 years and may have accumulated in the oceans, a hypothesis has been presented to suggest oceanwater transport as a possible pathway explaining the presence of PFOA in the Canadian Arctic. Currently there is insufficient data to evaluate the significance of this potential pathway .
Several researchers have indicate d that the timelines involved withtransport via ocean currents could not account for what appears to be rapidly increasing levels of perfluorinated substances in certain Arctic biota (Smithwick et al. 2006). While PFOA has been detected in co astal water and seawater even in remote areas (see section “2.2.8 Environmental Monitoring”, Caliebe et al. 2004, Yamashita et al. 2005), the extent to which this ma y be due to ocean or atmospheric transport is uncertain. Ocean water transport of perfluorocarboxycyclic compounds is a combination of :a) discharges of PFCAs to surface waters and transport to oceans; b) atmospheric loadings of PFCAs to surface waters and transport to oceans; and c) disc harge of precursors to surface waters, transformation to PFCAs and transport to o ceans (Prevedouros et al. 2006).

39OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

In addition to the possible role of aquatic transport via oceans to the Arctic, the possibility of atmospheric transport of PFOA on marine aerosols has been proposed (Prevedouros et al., 2006). Due to its nature as surfactant, PFOA is expected to be enriched on the water surface. As hypothesized, marine aerosols may be generated from this PF OA enriched surface through gas-bubble production and collapse through breaking waves and rough s ea conditions. The sea surface micro-layer may thus, supply the atmosphere with PFOA-rich partic les which undergo atmosphe ric transport over, at least, short distances. Studies are needed to determine whether and to what extent marine aerosols contain PFOA and contribute to their global transport. The dete rmination of whether perfluorocarboxylic acids are present, and to wh at extent, in marine aerosols, and whether this contributes to their global transport, is the s ubject of ongoing scientific investig ations (Prevedouros et al. 2006).

Conclusion Pure PFOA at room temperature has moderate vapor pressure (2.3 Pa). The vapor pressure of APFO is much lower with 0.008 Pa. APFO or PFOA disso lved in water dissociate to ions. Although the dissociated fraction is not subjec t to volatilization, depe nding on the pH, pure PFOA is expected to volatize from water to a certain degree. Due to emissions for more than 50 years, PFOA is distributed worldwide in the marine environment, and hence may be transported to remote areas via the aqueous pha se and the atmospheric phase. However, the significance of these source s are not currently known. Both atmospheric and aquatic transport mechanisms are actively being investigated. PFOA and PFOA precursors including fluorotelomer alcohols, olefins and perfluoroalkyl sulfonyl derivates are subject to long range transport. The relative environmental significance of these sources are not known currently.

2.2.5 Biodegradation A number of studies have examined the biodegrad ation potential of PFOA inaerobic and anaerobic conditions. Data have shown that PFOA is a highl y persistent substance unlikely to undergo rapid biodegradation in the environment. Slow degradation may be possible under certain conditions. Using an acclimated sludge inoculum, the biodegr adation of APFO was inve stigated using a shake culture study modeled after the Soap and Detergen t Association's presumptiv e test for degradation (3M, 1978). Both thin-layer and liquid chromatogra phy did not detect the pres ence of any metabolic products over the course of 2 1/2 months indi cating that PFOA does not readily undergo biodegradation. In a related st udy, 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 wa s conducted with a HPLC/MSD system. Although the results were deemed unreliable due to a lack ofdescription of experimental protocols or indications of a high degree of experimental error, several other studies c onducted between 1977-1987 also did not observe APFO biodegradation (3 M, 1979; 3M, 1980b; 3M, 1985; Pace Analytical, 1987). In addition, a study conducted by Oakes et al. (2004) indicated little biotic or abiotic degradation of PFOA on a time scale of 35 days, i.e., th e PFOA exposure concentrations were stable over time and ranged from 84.5 % to 114.5 % of the initial concentrations. Schroeder (2003), and Meesters and Schroeder (2004) investigated the biochemical degradation of PFOA in sewage sludge in laboratory scale reactors. After 25 days under aerobic conditions PFOA (initial concentration 5 mg/l) was not eliminated by metabolic processes, mineralization processes or by adsorption. In a recent study, Wang et al . (2005b) (see review in detail on page 27 ) report the biological degradation of the fluor inated alkyl group of radiolabel ed PFOA (< 0.6% in 28 days).

40OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Conclusion PFOA is not expected to undergo biodegradation.

2.2.6 Bioaccumulation

Bioconcentration is the process by which a chemical enters an organism and/or is adsorbed on to it as a result of exposure to the chemical in wate r – it often refers to a condition usually achieved under laboratory and steady state conditions. Bioaccumulation is the process by which a chemical enters an organism as a result of uptake through all possible routes of expos ure (dietary, dermal, and respiratory). Biomagnification is the process in which the chemical concentration in an organism achieves a level that exceeds that in the organism’s diet, due to dietary absorption

. Trophic biomagnification is the phenomenon where chemical concen trations in organisms increase with trophic level resulting in incrementally higher concentrations in upper trophic levels predators than in lower trophic levels.(Gobas et al. 2000) PFOA does not behave like lipophilic compounds thataccumulate in fat tissues. For lipophilic substances, accumulation is expected preferentially in the fat tissues. Due to the perfluorination, the hydrocarbon chains are ol eophilic and hydrophobic and the perfluorinated ch ains are both oleophobic and hydrophobic. In addition, functional groups attached to the perfluorinated chain (e.g., a charged moiety such as an hydroxyl group or sulfoni c acid) can impart hydrophili city to part of the molecule. Hydrophobicity is unlikely to be the sole driving force for the partitioning of perfluorinated substances to tissues because the oleophobic repellency opposes this partitioning process. Perfluorinated substances are also intrinsically polarchemicals because fluorine, a highly electronegative element, imparts polarity. Thus, perfluorinatedsubstances have combined properties of oleophobicity, hydrophobicity, and hydrophilicity over portions of a particular molecule. Due to these properties, the assumption that the traditiona l hydrophobic and lipophilic inte ractions between compound and substrate are the main mechanisms gove rning partitioning may not be applicable for PFOA.
To determine bioconcentration of PFOA, rainbow trout were expos ed in a flow-through system for 12 days followed by a depuration time of 33 days infresh water to determine tissue distribution and bioconcentration (Martin et al., 2003b). For determination of bioc oncentration, juvenile fish (5- 10 g) were exposed to a concentr ation of 1.5 μg/l in a flow-through system. At 7 occasions during uptake period and 9 occasions durin g depuration phase, fish were remo ved to determine the kinetics of uptake and depuration. Additionally, for the tissu e distribution study, four immature trout (30- 48 g) were exposed in separate tanks but under the same uptake conditions. PFOA concentration was highest in blood, kidney, liver and gall bladder and low in the gonads, adipose and muscle tissue. Within the blood, th e plasma contained between 94 – 99% of PFOA, with only a minor fraction detectable in the cellular fraction. Recove ry from hearts and spleen was low (<10%). A steady state was reached during uptake time. Visu al observation of depura tion data indicated possible biphasic depuration in blood, liver and carcass. However, this could not be verified statistically because of the small sample si ze. The following BCFs are calculated:

BCF carcass = 4.0 (+- 0.6); depurati on half-life: 5.2 d ( ± 0.67) BCF blood= 27 (+- 9.7); depurat ion half-life: 4.5 d ( ± 1.6) BCF liver = 8.0 (+- 0.59); depurat ion half-life: 3.9 d ( ± 0.28) PFOA occurs mainly in muscle, blood and organs (l iver, kidney) but not in lipid tissue and is reported for other species such as birds and mammals by several authors (see Table 20)

41OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Fathead minnows ( Pimephales promelas ) were exposed to PFOA in a static system to a concentration of 25 mg/L for 13 days, followed by a depura tion phase of 15 days. A BCF of 1.8 was calculated (3M, 1995). Daikin (2000) performed a bioaccumulation testaccording to OECD Guideline 305, with the carp Cyprinus carpio . The fish were exposed to PFOA concentr ations of 5 and 50 μg/l for 28 days. For the higher concentration of 50 μg /l, the steady state was reached af ter 16 days and a BCF of 3.2 was calculated. For the lower concentra tion of 5 μg/l, a BCF of 9.4 was determined after 16 days; this level was reduced to ≤ 5.1 after 28 days. No steady state was reached until end of exposition. Although experiments with fish and other aquatic sp ecies provide evidence that PFOA is not highly bioaccumulative, these results should not be extrapol ated to other animals. Fish gills may provide an additional mode of elimination and uptake whic h birds, terrestrial organisms, and marine mammals do not possess (Kelly et al., 2004). Martin et al. (2003a) exposed juvenile rainbow trout ( Oncorhynchus mykiss ) for 34 d to PFOA in the diet, followed by a 41 day depuration period. During the uptake period, animals were daily fed with spiked food (0.42 mg PFOA/kg food) at a rate of 1.5 % food pe r fish. Assimilation efficiency (% of PFOA absorbed relative to the amount fed) was 59 %, indi cating efficient absorption from food. At 6 occasions during uptake period and during depuration period, fish were removed to determine the kinetics of uptake and depuration. Carca ss and liver concentrations were determined by using liquid chromatography-tandem mass spectrometry, and kinetic rates were calculated to determine bioaccumulation parameters. The carcass uptake curves clearly showed by visual in spection, that the slope of the curve levels off by the end of the uptake period. According to the au thors the steady state was reached after 10 days. A depuration half-l ife time of 3d ( ± 0.42) and a BAF (Bioaccumulation factor) of 0.038 ( ± 0.0062) were determined. Martin et al. (2004a) examined PFO A contents in the food web from Lake Ontario (Canada). Adult lake trouts (top predator) were collected at various years and lo cations in Lake Ontario. Samples of prey fish (sculpins, smelts and alewifes) and macroinvertebrates (Mysis sp., Diporeia sp.) were collected at one location in October 2002. Lake trou t samples analyzed represented individual whole fish homogenates. The other species were processe d as composites of whole individuals. The mean PFOA content in Diporeia sp. and sculpin was 90 ng/g and 44 ng/ g, respectively. In the other fish samples contents of 1.0 to 2.0 ng/l and in Mysis sp. of 2.5 ng/g could be de tected. The authors note that Diporeia sp. is a benthic invertebrate species, and sculpins feed mainly in the benthic environment. Benthic contamination may therefore be th e source of contamination of this food web. As PFOA content in predators is lower than in prey species trophic biomagni fication of PFOA in the food web of Lake Ontario is unlikely occur. Trophic transfer of PFOA and othe r related perfluorinated compoun ds was examined in a Great Lakes benthic foodweb including water – algae – zebra mussel – round goby – smallmouth bass. In addition, perfluorinated compounds were measured inlivers and eggs of Chinook salmon and lake whitefish, in muscle tissue of carp, and in eggs ofbrown trout. Similarly, green frog livers, snapping turtle plasma, mink livers, and bald eagle tissues were analyzed to determine concentrations in higher trophic-level organisms in the food chain. Bi otic samples were collect ed from several rivers in Michigan and in the Calumet River in Indian a, USA. PFOA-concentrations in two of the sampling sites, Raisin River and St. Clair River, were 14.7 and 4.5 ng/l, respectively. The concentrations of PFOA in all tissue samples were above de tection limit but below the LOQ (see Table 20 Environmental levels of PFOA). Therefore, biomagnification of PFOA in the Great Lakes benthic foodweb is unlikely occur (Kannan et al., 2005).

42OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

The bioaccumulation of PFOA in the wild turtles Trachemys scripta elegans and Cinemy reevesii was reported by Morikawa et al. 2005. Serum con centrations of PFOA from 94 turtles were compared to surface water samples from the site of the turtle capture for several rivers in Japan. In Ai River water concentrations up to 87,100 ng/l were reported. Serum concentra tions in turtles collected in Ai River ranged from 47.1 to 115.6 ng/l, the corresponding BCF serum values ranged from 0.9 to 2.9. In Taisyo River water concentrations of 42.3 and 63.4 ng/l (two samples) and 9800 ng/l (one sample) were detected. Serum concentrations of 0.4 and 1.0 ng/l were reported for the turtles collected in low water concentra tion sides, and 7.6 ng/l were reporte d for turtles collected in high water concentration sides; corresponding BCF serum of 10-15.8 and 0.8 to 15.8 were reported with surface water concentrations ranging from 21.8 to 87,100 ng/l. However, as the wild turtles’ exposure to PFOA was probably not lim ited to surface water only, the BCFs reported by Morikawa et al (2005) may actually be BAFs. The biomagnification of PFOA and other perfluor oalkyl compounds in the food web of the bottlenose dolphin Tursiops truncates was investigated by Houde et al . (2006). Marine water, surface sediment, atlantic croaker ( Micropogonias undulatus ), pinfish ( Lagodon rhomboides ), red drum ( Sciaenops ocellatus ), spotfish ( Leiostomus xanthurus ), spotted seatrout (Cynoscion nebulosus ), striped mullet (Mugil cephalus ) and bottlenose dolphin samples were collected around the Charleston Harbor area (South Carolina). Marine wate r, surface sediment, zooplankton, sheephead (Archosargus probatocephalus ), pigfish ( Orthopristis chrysoptera ), pinfish, striped mullet, spotted seatrout and bottlenose dolphin samp les were collected at Sarasota Bay (Florida). At both locations seawater and surface sediment samples were collected in 2004. Fish samples were collected in 2002 and 2003 at Charleston and 2004 at Sa rasota Bay. Zooplankton samples were collected in Sarasota Bay in 2004. Samples of plasma, skin, and teeth we re collected from both locations during capturerelease events that took place in the summer of 2004. Additional dolphin tissue samples (i.e., liver, kidney, muscle and other) were collected during th e necropsies of recently deceased bottlenose dolphins from Sarasota Bay in 2002 and Charleston in 2003 (n=1 at each site). Also liver and kidney samples collected from stranded bottlenose dol phins were available at Sarasota Bay.
PFOA concentration in water was reported as 9.5 (±13) and 3.6 (±9.2) ng/l for Charleston and Sarasota Bay, respectively. Whole body c oncentration in fish ranged fr om <0.5 to 1.8 (±3.2) ng/g for Charleston and <0.5 to 0.3 (±0.5) for Sarasota Ba y. PFOA concentration in the plasma of dolphin was 43 (±24) and 3.4 (±3.5) ng/g for Charle ston and Sarasota Bay, respectively. The authors calculated biomagnifica tion factors for several predators and their prey, but it is difficult to understand how they calculated thes e factors. For example: At Charleston, for seatrout/pinfish a BMF of 7.2 is reported. However, a recalculation from data given in the report, indicates seatrout residues (1.8 ng/g) divided by pinfish residues (<0.5 ng/g) should be 3.6. The BMFs for dolphins and their prey can not be re calculated because they are based on whole body concentration in dolphins. The ex trapolation factors from plasma and tissue to whole body is not transparent. In addition the water and biota samples were originating from different years with high variabilities. Keeping these un certainties in mind, the maximu m BMF for dolphinwhole/striped mulletwhole and dolphinwhole/pinfishwhole was 13. As this BMF is >1, it suggests a potential for biomagnification in this pr edator-prey relationship. An eastern Arctic food web was analyzed for PF OA and other perfluorinated compounds (Tomy et al., 2004). Liver tissues of beluga whale ( Delphinapterus leucas ), narwhal (Monodon monoceros ), walrus ( Odobenus rosmarus ), deepwater redfish ( Sebastes mentella ), glaucous gullus ( Larus hyperboreus ), and black-legged kittiwake ( Rissa tridactyla ) and whole organism homogenates of Arctic cod ( Boreogadus saida ), shrimp ( Pandalus borealis; Hymenodora glacialis ), clams (Mya truncata ; Serripes groenlandica ), and mixed zooplankton were analyze d. PFOA concentrations in clam and black-legged kittiwake samples ( n =5) analyzed were below Method Detection Limits (MDLs) of 0.2 ng/g. Mean concentrations of PFOA in zooplankton ( n =5) were 2.6 ± 0.3 ng/g. PFOA was de43OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

tected in 3 of 7 shrimp samples at a mean c oncentration of 0.17 ± 0.06 ng/g. PFOA was detected in only a single Arctic cod (0.47 ng/g) ( n =5). PFOA was detected in only 2 of 5 the samples of deepwater redfish at a mean concentration of 1.2 ± 0.8 ng/g. Statistical differences in the mean PFOA concentrations were observed between walrus (0.34 ± 0.09 ng/g) and narwhal (0.9 ± 0.1 ng/g; p < 0.005) and between walrus an d beluga (1.6 ± 0.3 ng/g; p < 0.05). Biomagnification of PFOA between individual feeding relati onships can be seen but not through the entire food web: predator prey BMFs walrus clam 1.8 narwhal arctic cod 1.6 beluga whale arctic cod 2.7 beluga whale redfish 0.8 black-legged kittiwake arctic cod 0.3 glaucous gullus arctic cod 0.6 arctic cod zooplankton 0.04

Conclusion The results indicate a low potenti al for bioaccumulation of PFOA in fish and low biomagnification potential in several aquatic food webs. In some marine and Canadian Arctic mammalian food web studies, a potential for bioma gnification has been suggested.

In tests with the rainbow trout Oncorhynchus mykiss a bioconcentration f actor (BCF) of 0.038 and bioaccumulation factors (BAF) for organs of 27 (b lood), 8.0 (liver) and 4.0 (carcass) were obtained. These laboratory studies indicate a low bioaccumulation potential in fish. Some monitoring data suggest a low biomagnification poten tial in aquatic food webs, while in some marine and Canadian Arctic mammalian food web studies a potential for biomagnification has been suggested. Further elucidation of the mechanisms leading to uptake and accumulation in biota is required. For substances like PFOA, bioconcentration may not be the most relevant e ndpoint to consider. For PFOA and other perfluorinated subs tances, biomagnification appears to be occurring in terrestrial and marine mammals. An assumption of the BAF/BCF approach is that bioaccumulation occurs by the same mechanisms for all chem icals in both water-b reathing animals (e.g., fish and aquatic invertebrates) and air-breathing animal s (e.g., terrestrial mammals, birds and marine mammals) resulting in a similar bioaccumulation poten tial between these organism cla sses for a particular substance. Both Kelly et al. (2004) and Mackay and Fraser (2000) discuss key differences in the bioaccumulation mechanisms between air-breathing animals and water-breathing animals. Furthermore, Kelly et al. (2004), describe how th e current BAF/BCF/log K owparadigm may not be adequate to assessbioaccumulation potential in air-breathing animals. From a scientific perspective, additional measures of bioaccumulation which also address the pot ential for chemicals to biomagnify include biomagnification factors and trophic magnifi cation factors (Gray 2002; MacKintosh et al. , 2004; Martin et al. , 2004a; Tomy et al. , 2004; and Cabana and Rasmussen, 1994).

2.2.7 Other Information on Environmental Fate

Behaviour in soil The adsorption-desorption of APFO was studied in 25 ml solutions of 14 C-labeled APFO in distilled water with 5 g Brill sandy loam soil for 24 hours at a temperature of 16-19 oC. The study reported a Kd of 0.21 and a Koc of 14 indicating that PF OA has high mobility in Brill sandy loam soil (3M 1978b). The K OC value, however, is questionable due to the lack of accurate information on the purity of the 14 C-labeled test substance (Boyd, 1993a,b).

44OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Moody and Field (1999) conducted sa mpling and analysis of sample s 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. Perf luorooctanoate was detected at maximum concentrations ranging from 116 to 6750 μg/l at the two sites many years after its use at those sites had been discontinued. These results suggest that PFOA can leach to groundwater. An adsorption-desorption test according to OE CD guideline 106 was made by APME (2003) at DuPont, Newark sponsored by Plasti cs Europe. APFO was tested with four soil and one activated sludge samples (equilibration time 24 h). Quantification (analytics: LC-MS/ MS) was made using a calibration curve. The K OM values ranged from 28 l/kg to 133 l/kg. The data from these batch sorption studies are of limited utility in understanding the movement of PFOA released to soil. Batch sorption studies, because of their limited nature, do not provide all the information needed to understand the be havior of PFOA in the environment. Extensive site specific monitoring of soil and ground water concentrations of PFOA and related substances was conducted by 3M, DuPont Daikin and others. PFOA in soil has been shown to persist for decades and to be a long term source of groundwater and surface wa ter contamination (see for example 3M (2005a), DuPont (2003a)). At the DuPont Washington Works site soil cont aminated by perfluorochemical waste has been shown to contain ppm levels of PFOA 3 decades after application ceased. The underlying groundwater also contains ppm levels of PFOA (DuPont, 1999) Extensive field monitoring data generated by 3M at the Decatur, AL site have have also shown that PFOA is persistent in soils. Soil samples were co llected from a former sludge application area of the 3M Decatur, AL facility also show soil contamination and underlying groundwater contamination up to ppm levels decades af ter application ceased.
Conclusion PFOA is not rapidly degraded in soil. PFOA has a ve ry low to weak adsorption to soil organic matter (classification according to Blume and Ahlsdorf (1993) witha soil sorption coefficient Koc between 1 and 100). Therefore a high mobility of PFO A in soils can be assumed. Monitoring data show that PFOA persists in soil and leaches over tim e. PFOA in soil can be a long term source to underlying groundwater

2.2.8 Environmental Monitoring Measurements related to Italy, Tri ssino (Miteni), APFO production site According to Miteni, information about measurements is not available.

Measurements related to The Netherlands , Dordrecht (DuPont), APFO use site According to DuPont, no information about measurement of PFOA is available from monitoring of surface water and biota. Measurements related to Germany, Gendorf (Dyneon and Clariant), use and former production site Dyneon conducted environmental monitoring meas urements in and around the Gendorf site in 2001/2002. Another environmental monitoring campai gn with lower detection limits will be completed in 2006. The new sampling campaign includes soil and plants, in a ddition to the water and fish measurements performed in 2001/2002.

45OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Water Groundwater wells were sampled in 2002 (locati ons upstream, under, and downstream of plant site). All measurements gave results be low the limit of detection (LOD) (200 μg/l). All measurements in the receiving stream, river Alz, sampled in 2001 showed results below LOD. The LOD for water was 200 μg/l, for sediment 300 μg/kg and for suspended solids 600 μg/l.

Biota The sampling results from river Alz biota are described in Table 19.

Table 19 Concentration in Fish sampled at Alz river, in 2001[μg/kg wet weight]

Eel (n=4) Barbel (n=3) Chub (n=3) At WWTP outfall, liver 500, 600, 1600, 3200 < LOD 500, 2× < LOD At WWTP outfall, filet 2 ×< LOD, 1800, 2400 < LOD < LOD

Upstream of plant, liver & filet < LOD < LOD < LOD

LOD: 500 μg/kg

Measurements related to USA, Alab ama, Decatur (3M), APFO use site Environmental monitoring data on the 3M Decatur, AL facility available to date are summarized in this document. More data are being generated and will be made available to the public through the US EPA Docket in 2007.

Soil Soil samples were collected from the former sludge application area, former ly investigated areas and background areas of the 3M Decatur, AL f acility during February and March, 2006. Samples were collected at 0 - 3 in., 3 - 6 in., 6 - 12 i n., several 0.5 foot intervals between one and ten feet and several five foot intervals be tween ten feet and selected lowe r competent layers. The range of PFOA concentrations for the sl udge application fields was 29 -3140 ppb for 0 - 3 in.; 34.2-5690 ppb for 3 - 6 in.; 18.5 - 1870 ppb for 6 - 12 in.; 25.3 - 8780 ppb for 1 - 10 feet; and 14.1-3800 ppb for over 10 feet. Background PFOA levels in the sludge application control and in the non-sludge application background area ranged between 0.2 (non-detect ) and 9.92 ppb. The soil samples collected from the borings in the formerly investigated areas had a range of PFOA levels between 0.2 (nondetect) and 810 ppb (3M, 2005b).

SedimentSediment samples (12) were collected in the Tennessee River adjacent to the 3M Decatur, AL facility in April, 2006 and analyzed fo r PFOA concentrations. The range of concentrations of PFOA for 12 sediment samples is between 0.393 and 24.1 ng/g (ppb) (3M, 2005b). Wastewater treatment effluent All process wastewaters from the 3M and Dyne on manufacturing operations are treated in this treatment facility which include s both physical-chemical and biol ogical treatment. Water samples were collected prior to discharg e to Baker’s Creek which then em pties into the Tennessee River. This water includes both the wastewater treatmen t effluent as well as non-contact cooling water

46OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

used in the manufacturing process. PFOA was present at an average of 602, 766, 1028, 310, 58, and 88.3 μg/l for samples collected in 1998, 1999, 2000, 2001, January 2003, and May 2003, respectively (3M, 2003b; Santoro, 2003; AR226-1482).

Groundwater
Groundwater concentrations of PF OA were measured in the former sludge application area, the sludge application control and the ar ea wells at the 3M Decatur, AL facility in February and March, 2006. Wells in the sludge application area co ntained PFOA in the range between 2.27 and 3857 μg/l. The control sludge application wells contained PFOA between 0.220 – 94.7 μg/l. The LOI PFOA levels ranged from 0.025 (non- detect) and1333 μg/l (3M, 2005b). Surface water Surface water samples (12) were collected in the Tennessee River and a nearby creek called Baker’s Creek adjacent to the 3M Decatur, AL facility inApril, 2006. Each sample was analyzed twice for PFOA. Other surface water samples (16) were taken in a longitudinal direction along the length of the river along with a set of 30 samples collected across the river. The range of averaged concentrations of PFOA for the 12 samples is 0.0734 to 0.420 ng/ml (ppb). The range of PFOA concentrations for the longitudinal samples (16) is 0.00492 (non-detect) -0.0260 ng/ml. The range of PFOA concentrations for the transect samp les (30) is 0.00266-0.435 ng/ml ((3M, 2005b).

PlantsThe 3M Decatur, AL facility collected vegetation in its former sludge application area, sludge application control area and background locations on October 7, 2004. Samples were taken of four vegetative species found in the sludge application areas; br oomsedge, pokeweed, curly dock, and goldenrod. PFOA concentrations in the vegetation in the former sludge application fields ranged from 42.3 – 912 ng/g (ppb). In the sludge control and the background area the PFOA concentrations ranged from 1.99 – 9.22 ng/g (ppb) (3M, 2005b).

Invertebrates
Asiatic clams were collected from the same locations as freshwater fish at the 3M Decatur, AL site in December, 2004. PFOA concentrations in clam s upriver were non-quantifiable and downriver levels of PFOA ranged from 0.508 ng/g to 1.01 ng/g at Baker’s Creek (3M, 2005b). Freshwater fish Two species of freshwater fish, channel catfish and large mouth bass were collected from the Tennessee River and Creeks in the vicinity of the 3M Decatur, AL facility during December, 2004. Filet and whole body parts of both species were analy zed for PFOA. The upriver PFOA levels for filet and whole body ranged between no n-detect and 5.88 ng/g. Downriver samples of filet and whole body ranged from non-quantifiable to 6.06 ng/g. The 6 ng/g sample was found in a large mouth bass in Baker’s Creek adjacent to the 3M facility, a creek that empties into the Tennessee River (3M, 2005b).

Mammals Small mammal sampling was conducted concurrentlywith vegetation sampling from October 5-7, 2004 at the 3M Decatur, AL facility. Hispid cotton ra t were collected in the same areas as the sampled vegetation. Livers and serum were collected from each rat and analyzed for PFOA. PFOA was present above the 400 ng/ml detectio n limit in two of the fifteen sampled rat serum samples but below the quantitation limit of 1000 ng/ml. All liver samples meas ured non-detect for PFOA (3M, 2005b).

47OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Other Media Surface water samples (12) were collected in the Tennessee River adjacent to the 3M Decatur, ALfacility in April, 2006. Each sample was analyzed twice for PFOA. The range of averaged concentrations of PFOA for the 12 samples is 0.0977 to 70.4 ng/ml (ppb) (3M, 2005b). Measurements related to USA, Alabama, Guin (3M) Surface Water Water samples were collected in September 2005 from 2 locations within the 3M Guin facility. Water collected from a surface water location contai ned concentrations of PFOA ranging from <LOQ (<0.0246 ng/ml) to 10.8 ng/ml (3M, 2005a). Water samp les collected from a municipal water pipeline within the 3M Guin facility contained PFOA at concentrations of <LOQ to 9.1 ng/ml (3M, 2005a). Measurements related to 3M’s Cottage Grove, MN (USA)

Soil Soil samples beneath a “ponded” area at the land fill contained PFOA at concentrations of 22.3 and 31.1 ng/g. Concentrations of PFOA were measur ed in soil borings at the landfill groundwater pumpout/spray irrigation site at depths from 0 to 26 feet below ground surface PFOA in the range from <0.208 to 21.6 ppb. PFOA concentrations in a background soil ranged from 1.28 to 4.87 ppb. (Oliaei et al., 2006).

Sediment Sediment cores (collected in 2005 ) obtained at a depth of 10 cm from the river bottom contained PFOA at 18, 6.62, 1.31, and <0.301 ppb for samples coll ected at the 3M wastewater treatment discharge site, and 3 sites located progressively downstream, respectively (Oliaei et al., 2006). Wastewater Treatment ef fluent and sewage sludge The MCES wastewater treatment plant in St. Paul, MN receives leachate from the Pine Bend landfill and PFC-containing groundwater from the Oakda le dump. Influent and effluent samples collected in 2005 contained PFOA at concentratio ns of 46 and 78 ppt, respectively. Primary sludge, secondary sludge, and biosolid samples co ntained PFOA at 3.79, 21.5, and 11.1 ppb, respectively (Oliaei et al., 2006). The 3M wastewater treatment plant at Cottage Grove, MN discharges from the plant to the Mississippi Ri ver. Concentrations of PFOA in efflue nt of the plant ranged from a high of 1,991 μg/l in January/March 2000 (to a low of 17.7 μg/l in June 2003. (3M, 2003b; Santoro 2003 – AR 226-1482). Influent and effluent PFOA concentra tions from this facility in 2005 (single sampling date) were 3.74 and 62.4 μg/l , respectively (Oliaei et al., 2006). The increase in PFOA concentration has been previously attributed to th e degradation of PFC precursor compounds to PFOA during the treatment process (Schulz et al., 2005). Landfill Leachate and gas
Pine Bend landfill received wastew ater sludges from the 3M Cottage Grove wastewater treatment plant beginning in 1975. Leachate samples collected in 2005 from the combined leachate tanks contained PFOA ranging from 14.2 ng/ml to 81.8 ng/ml while a gas condensate sample from the landfill’s active gas collection system containe d PFOA at 83.8 μg/l. (Oliaei et al., 2006) .

48OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Groundwater 3M disposed of fluorochemical wastes from the Cottage Grove site at the unlined Washington county landfill from 1969-1975. Groundwater sample s collected in 2004 upgradient and downgradient from the landfill contained PFOA at <0.0 027 and 41.6 μg/l, respectively. Groundwater samples collected in 2005 upgradient and downgradient from the Pine Bend landfill contained PFOA at concentrations of 0.008 and 1.6 μg/l, respectively (Oliaei et al., 2006). Surface water Surface water samples (collected in 2005) from th e 3M Cottage Grove wastewater treatment plant discharge site water, the discharge site river wa ter upper layer and from 3 sites located progressively downstream from this site cont ained PFOA at 3250, 3650, 35.30, <4.91 and <4.90 ppb, respectively (Oliaei et al., 2006).

Plants Samples of duckweed collected in 2005 at the site receiving the Cottage Grove wastewater treatment discharge contained PFOA at18.10 ppb. (Oliaei et al., 2006)

Freshwater Fish
Three sets of freshwater fish were collected in the vicinity of the 3M Cottag e Grove facility in 2004 and 2005. Twenty fish were collected in August 2004 from the immediate vici nity of the 3M Cottage Grove wastewater treatment discharge area. Concentrations of PFOA in liver samples from these fish ranged from <0.359 to 0.551 ppb in sma ll mouth bass, <0.374 to 1.27 ppb in white bass, <0.345 to 0.662 ppb in common carp, and <0.357 to 6.51 ppb in walleye. Over 100 fish were collected in October 2005 in a pool downstream and proximate to the 3M Cottage Grove discharge location. Carp blood from this lo cation contained PFOA at concen trations of 0.87 to 15.5 ppb. Blood samples from other fish species generally contai ned low to non-detectable concentrations of PFOA (Oliaei et al., 2006). Drinking water PFOA was measured in water samples collected from the Oakdale municipal well at concentrations ranging from not detected to 0.858 μg/l in December 2004 (2/12 samp les below detection limit of 0.025 μg/l) and 0.288 to 0.856 μg/l in August 2005 (0/8 samples below detection limit of 0.025 μg/l) (Bilott, 2005). Measurements related to DuPont’s Chambers Works, Deepwater, NJ (USA)

Wastewater Treatment Plant
Influent concentrations of PFOA to the Chambers Works wastewater treatment plant were 194 and 26 μg/l based on two samples collected on April 21 and May 15, 2003, respectively. Effluent concentrations of PFOA directly from the wastew ater treatment plant on April 18 and May 16, 2003 were 475 and 279 μg/l, respectively. Concentrati ons of PFOA discharged from the wastewater treatment plant following the addition of non-contact cooling water to the Delaware river were 193 and 102 μg/l on April 18 and May 16, 2003, respectiv ely. Sampling at this location on June 10/11, 2003 resulted in PFOA concentrations in wa ter of 76, 144 and 97 μg/l (EPA, 2004b – AR226- 2620).

Groundwater Groundwater from 6 monitoring wells located on th e Chambers Works site contained PFOA at concentrations ranging from 191 to 5230 ng/l (all samples were positive; LOQ is 0.050 μg/l; sampling

49OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

in May/June 2003) (EPA, 2004b – AR226-2620). Five in terceptor wells contained PFOA at concentrations of 354 to 46,800 ng/l (a ll samples were positive; sa mpling in May 2003) (EPA, 2004b – AR226-2620). Surface water
Concentrations of PFOA in Delaware river wate r samples collected in June 2003 from 30 locations ranged from not quantifiable (LOD of 0.010 μg/l , LOQ of 0.050 μg/l) to 0.577 μg/l (EPA, 2004b – AR226-2620). Measurements related to DuPont Washington Works site (USA)Environmental monitoring data on the DuPont, Parkersburg WV facility available to date are summarized in this document. More data are being generated and will be made available to the public through the US EPA Docket in 2006 and 2007.

Air Air samples were collected with OSHA Versatil e Sampling (OVS) method tubes and High-Volume Sampling (HVS) equipment within a two mile radi us of the Washington Works, DuPont site in West Virginia from August-October, 2005. The OVS samples ranged from limit of detection (0.09- 0.64 ng/tube fraction) to 60.2 ng/m 3 of PFOA. HVS samplers measur ed PFOA in the range of 0.01 to 75.9 ng/ m 3 . (EPA-HQ-OPPT-0113-0032). Air samples were collected during a ten week period from November 2003-January 2004 along the fence lin e of the DuPont Washi ngton Works plant in West Virginia. OVS tube monitoring during the ten weeks measured PFOA levels from <120 ng/m 3 to 900 ng/ m 3 (Barton et al., 2006).

Soil Soil samples were collected on-site at the DuPont Washington Works West Virginia site from soil 3 - 6 in. and 6 - 12 in. below the surface. The so il samples collected from the 3 - 6 in. intervals ranged from 72- 130 ppb PFOA. Those soil samples in the 6 - 12 in. interval contained PFOA in the range of 44-130 ppb PFOA (EPA-HQ-OPPT-0113- DuPont Quarterly MOU Status Report #2 Text, Tables and Figures). Soil samples below the site of former waste management units and anaerobic digestion ponds showed PFOA concentrations ranging from 18 to 48000 ug/kg. The highest levels were from 8-12 foot deep samples (DuPont, 1999).

Groundwater Groundwater has been monitored at the Washingt on Works facility in West Virginia from 1996- 2006 at groundwater monitoring well locations on the plant site. APFO levels have been measured since 1996 and PFOA levels have been measured since 2004. APFO levels at 70 different groundwater wells at the plant between 1996 and 2006 ranged from < 0.050 to 4.66 x 10 5 ug/l. PFOA levels in groundwater at Wa shington Works between 2004 and 2006 ranged from 0.537-118 ug/l (EPA-HQ-OPPT-0113-0032). Groundwater monitoring wells located at three differe nt landfill sites in the vicinity of the DuPont, West Virginia Washington Works plant were monitored for PFOA and APFO between 1996 and 2006. The landfills are the Dry Run landfill, the Lo cal landfill and the Letart landfill. PFOA and APFO levels at the Local landfill ranged from 0.141-69 ug/l and <0.050-91.1 ug/l respectively. Forthe Dry Run landfill, PFOA and APFO ranged from <0.050-32.6 ug/l and <0.050-39.0 ug/l respectively. PFOA and APFO levels at the Letart la ndfill were measured at a range between 0.126- 20,200 ug/l and 0.058-57,200ug/l, respect ively EPA-HQ-OPPT-0113-0032).

50OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Groundwater samples below the site of former waste management units and anaerobic digestion ponds showed PFOA concentrations rangi ng from 0.1 to 13,600 ug/kg (DuPont, 1999).

Freshwater Fish Channel catfish and large mouth bass were collected from the Oh io River near the Washington Works, DuPont WVA site and analyzed for APFO and PFOA in filet, viscera and carcass. PFOA levels in viscera samples for both species ranged between 1.23 ng/g and 1440 ng/g; PFOA in carcass samples ranged from 0.412 to 10.8 ng/g; and th e filet samples were between 0.234 ng/g and 9.17 ng/g PFOA. Background fish had the following ranges of PFOA: viscera- 0.85-42.6 ng/g; carcass- 0.087-1.59 ng/g; filet- 0.146-1.65 ng/g (EPA -HQ-OPPT-0113-0032 DuPont Quarterly MOU Status Report #2 Text, Tables and Figures).

Plants Grass samples collected from on-site and off-site at the Washington Works DuPont West Virginiasite were analyzed for PFOA. Samples were also collected at a background location many kilometers away. PFOA was detected in all samples analyzed. On-site sampled were washed to help evaluate inthe PFOA was in the plant tissues or on the surface. On-site unwashed grass samples contai ned PFOA levels between 154 ppb and 574 ppb on a dry weight basis. Washed grass samples ha d PFOA levels between 33.5 ppb and 92.8 ppb (dry weight) (EPA-HQ-OPPT-0113-0032 DuPont Quarte rly MOU Status Report #2 Text, Tables and Figures). For off-site sampling DuPont collected of the a bove ground plant material and the underlying soil layer. In their report DuPont labeled these samples as 0 - 1 in. and 1 - 3 in. soil/grass samples. The 0 - 1 in. samples contained grass and the underlying thatch layer. The 1 - 6 feet samples contained the underlying soil and root material. Off-site grass samples were collected during a dry period and following a rainfall event. Grass was sampled from 14 sites within 3.2 km of the facil ity and a background site far from the facility. Analysis of vegetation and underlying soil from all location sampled had detectable PFOA. The background samples and samples from one location n ear the facility had levels below quantitation limits (<6.0 ppb dry period and 5.0 ppb wet period). The other 13 locations had PFOA concentrations on grass ranging 6.88 ppb - 118 ppb during th e dry period and < 7.0 – 121 ppb following the rain event. The underlying soil laye r material collected during the dry period between 1 - 6 in. in depth contained PFOA at <7.0 to 52.8 ppb. Underl ying soil following the rain event had PFOA concentrations ranging from <7 ppb to 51.6 ppb. A ll concentrations were on a dry weight basis (EPA-HQ-OPPT-0113-0032 DuPont Qu arterly MOU Status Report #2 Text, Tables and Figures).

Other Media DuPont has been sampling public water supply syst ems in West Virginia and Ohio for APFO since 2002 and for APFO and PFOA since 2004. Twelve system s are sampled in West Virginia and seven systems are sampled in Ohio each year for A PFO and PFOA. Additionally, DuPont has sampled twenty-five residential water wells in the vicini ty of Washington Works on a quarterly basis for APFO since 2002 and for APFO and PFOA since 2004. The PFOA concentrations in the public supply systems have ranged from a low of 0.001 ug/l (non-detect) to 71.9 ug/l or 71,900 ng/l. The residential well samples have ranged from a lo w of 0.05 ug/l (non-quantif iable) to 27.1 ug/l or 27,100 ng/l (EPA-HQ-OPPT-0113-0032 DuP ont Quarterly MOU Status Report #2 Text, Tables and Figures).

51OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Rainwater Precipitation samples were collected at the DuP ont Washington Works site, West Virginia August 25-26, 2005 at the same locations where ambient air samples were collected. Precipitation sampleshad PFOA levels that ranged from < 8.8 ng/l to 1660 ng/l. Out of eleven samples, seven were nonquantifiable, four ranged between 10.2 ng/l and 52.9 ng/l and one measured 1660 ng/l(EPA-HQOPPT-0113-0032 DuPont Quarterly MOU Status Report #2 Text, Tables and Figures).

Environmental Monitori ng Analytical Methods Yamashita et al. (2004a) investigated sources of perfluorooctanoic acid ( PFOA) contamination during the analytical pathway. They identified possibl e contamination caused either by instrumentation or the procedure applied. One source of procedur al contamination is fluoropolymers, such as poly(tetrafluoroethylene) and pe rfluoroalkyloxy compounds, which are present in a variety of laboratory products. Therefore environmental or biolog ical matrixes, samples or extracts are not allowed to come in contact with such fluoropolymers duri ng the analysis. The authors detected PFOA in injections of 10 μL of pesticide- grade methanol taken inan autosampler vial. The HPLC/MS/MS instrument (HPLC tubing, internal fluoropolymer pa rts and autosampler vial septum) was identified as the possible source of this contamination. PFOA was the most abundant compound, found at approximately 30 pg per 10 μ L of methanol injected. PFOA was al so detected in polypropylene sample bottles. After extraction with methanol, PFOA was the only target compound detected in polypropylene containers, and the highest concentration found was 27 ng/l. Furthermore, notable amounts of PFOA (46 pg/l) were detected in the Se p-pak cartridge. Three types of purified reagent water were examined for the occu rrence of perfluorinated carboxylic acids. Distilled water prepared using the Yamato distillation apparatus contained considerable amounts of target fluorochemicals. Milli-Q and HPLC-grade waters had relatively lower background levels. Based on measurements, the contamination from HPLC-grade water was es timated to be 1.2 pg/l of PFOA. After extraction, extracts often were filtered. Nylon syringe filters contained trace amounts of PFOA. Boulanger et al. (2004) and So et al. (2004) id entified fluoropolymer and fluoroelastomer containing devices as sources of PFOA contamination. Because this contamination cannot be omitted, poor detection limits or recovery rates significantly >100% were observed. Ellis et al. (2004), and De Silva and Mabury (2004) used GC/MS to avoid instrumental contamination caused by HPLC. However, the drawback of this procedure is the additional derivatization step. Studies which do not provide info rmation on procedural blanks, re coveries or detection limits should be treated with care.

The first worldwide interlaborat ory study of PFCs in human and environmental matrices (van Leeuwen et al. 2005) indicat es that analytical resu lts from individual laborat ories are not directly comparable due to variability in analytic al results between individual laboratories.

Air Only a few studies were available for the determin ation of PFOA in air. Reagen et al. (2004) developed an analytical procedure using OSHA versat ile sampler (OVS) tubes and LC/MS for the determination of PFOA at the work place. A limit of quantitation ( LOQ) of 8 μg/m 3 , a limit of detection (LOD) of 2.4 μg/m 3 , and a recovery rate of 94% were re ported for APFO. Kais er et al. (2005a) obtained a LOQ of 0.474 μg/m3 using electrospray LC /MS or LC/MS/MS and r ecoveries in the range of 88 and 95.7%. Moriwaki et al. (2003) determined PFOA in dust samples collected in Japanese homes. A LOQ of 50 ng/g and a rec overy rate of 73% are reported.

52OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Water In general, PFOA is analyzed in water samples (s urface-, drinking-, sea-, and waste water) using HPLC/MS/MS. Most methods applied have detection limits in the pg/l (e.g. Taniyasu et al. 2005a; Saito et al. 2004b; Berger et al. 2004; Tseng et al. 2006; Caliebe et al. 2004) to low ng/l range; the corresponding limits of quantification are somewhat higher (e.g. Saito et al. 2004; Berger et al. 2004). Scott et al. (2006b) developed a method for determining very low levels (detection to 0.01ng/l) of PFCAs in large volume samples. Good agreement was shown between this method and the common LC/MS/MS method for samples from sewage treatme nt plant effluent, Lake Superior samples and precipitation samples in remote sites in Canada. PFOA levels in sewage treatment plant effluent ranged between 27.4 and 34.0 ng/l for the XAD resin and 30.5 to 40.2 for the LC/M/MS method. Samples taken from 260 meters deep in Lake Superior measured 0.26 ng/l on the XAD resin and 0.35 ng/l with the LC/MS/MS method. Precipitation samples in the remote regions of Canada analyzed by GC/MS contained 3.1 and 1.6 ng/l of PFOA for Turkey Lakes and Kejimkujik respectively (Scott et al. 2006a). Sediment and sewage sludge Schroeder (2003) used LC/MS and flow injection analysis for th e detection of PFOA in sewage sludge (LOD: 10 mg/kg dw). Berger et al. (2004) and Kallenborn et al. (2004) achieved a LOD of 6 ng/g ww and a LOQ of 200 ng/g ww using HPLC/MS for analysis of sediments and sludges. De Voogt and van Roon achieved a LOD of 0.4 ng/g dw for soil samples via HPLC/MS. Higgins et al. (2005) developed a HPLC/MS/MS procedure for th e determination of PFOA in sediment and sludge samples. They obtained LODs of 0.011 ng/g dw (sediment) and 1 ng/ g dw (sludge); the corresponding recovery rates were 88 and 71%.

Biota
Most of the scientis ts used the ion-pair extraction procedure of Hansen et al. (2001) for the analysis of biota samples. In most cas es HPLC/MS/MS was used for m easuring PFOA concentrations. In some cases the LODs and LOQs were relatively hi gh (>30 ng/g ww) due to matrix effects (e.g. Sinclair et al., 2004; Kannan et al., 2002a; Van de Vijver et al. 2005). Ta niyasu et al. (2005a) and So et al. (2006a) developed very sensitiv e analytical methods for the dete rmination of PFOA in tissue of animals; PFOA concentration of 0.01 (=LOD) and 0.2 ng/g ww (LOQ) could be detected.

Other Media Using LC/MS/MS, Mawn et al. (2005) developed methods to quantify the amount of PFOA extracted from textiles such as apparel, home furnishings and carpet samples through their contact with water, methanol, and sweat and saliva simulants. The LOQ for samples extracted in water and sweat simulant is 1 ppb (ng PFOA/g sample) while the limits of quantitati on for samples extracted in saliva simulant and methanol were 3 ppb and 2.5 ppb, respectively. In SNF (2006) a HPLC/MS method is described for the determination of PFOA in all-weather jackets. The LOD is not reported; the lowest measured value was 6.1 ng/g. Larsen et al. (2005) determined perfluorooctano ic acid in polytetrafluor oethylene polymers using pressurized solvent, reflux ex traction methods and LC/MS/MS. The available analytical methods are compiled in the Annex.

53OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

PFOA in Environmental Sa mples, temporal trends Using current analytical methods PFOA might be detected in concen trations in th e range of ng/m 3 (air), pg/l (water), and pg/g (ww; biota samples) . Improvement of the analytical methods is ongoing. Today, PFOA is quantifiable in environmental matr ices where the previous analytical methods failed. Studies using improved analytical sensitiv ity result in greater de tection frequencies. Only a few studies examined PFOA levels in e nvironmental samples over time (Holmström et al., 2005; Lucaciu et al., 2005; Smithwick et al., 2006; 3M, 2003b). PFOA was not detected in any of the 146 guillemot egg samples collected from 1968 to 2003 in the Baltic Sea, so no trend could be determined. Over this same period, however, perf luorooctane sulfonate (PFOS) levels increased (1968-1997) and then decreased (1997-2003) (Holms tröm et al., 2005); in many cases PFOS and PFOA are simultaneously detected in environmenta l samples. Lucaciu et al. (2005) determined PFOA concentrations in suspended sediment samp les collected at Niagara-on-the-Lake from the Niagara river. They observed increasing PFOA con centrations from 1981 (ca. 80 pg/g) to 1995 (ca. 220 pg/g). Between 1996 and 2000, PFOA concentrations decreased to ca. 100 pg/g. In the samples taken in 2001 the highest PFOA concentration wa s measured (ca. 290 pg/g). Smithwick et al. (2006) analyzed archived polar be ar liver tissue samples from two geographic locations in the North American Arctic collected from 1972-2002. The eastern group (northern Baffin Island, Canada) comprised 31 samples, the western group (Barrow, Alaska) 27 samples. Between 1972-2002, in the eastern group a significant increase in PFOA concentration was observed (1972: ca. 3 ng/g ww PFOA; 2002: ca. 20 ng/g ww PFOA). Only slight changes in PFOA concentrati ons were detected in liver tissue samples collected at the western lo cation (1972: ca. 2 ng/g ww PFOA; 2002: ca. 5 ng/g ww PFOA). PFOA was measured in liver tissue from 80 adult female sea otters collected from the California coast. These otters were found freshl y dead and beached along the coast. During 1992 – 2002, concentrations of PFOA ranged from <5 to 147 ng.g -1 ww (Kannan et al ., 2006). Concentrations of PFOA also increased significantly from 1992 to 2002 for adult female sea otters. However, PFOA was not found in the adult male sea otters (detect ion limit of 5 ng.g -1 ). The reason for this lack of detection is not known. It should be noted that PFOA was found in the procedural blanks and in methanol injections performed between samples, however, as this background PFOA signal was consistent, it was subtracted from the ca libration curve and samp les (Kannan et al., 2006).Based on the results of 3M (2003b) wastewater effluent con centration collect ed at Decatur (Alabama, USA), the average c oncentrations significantly decr eased between 1998 (602,000 ng/l) and 2003 (88,300 ng/l).

Air Oliaei et al. (2006) sampled gas condensates at the Pine Bend Landfill, Inver Grove Heights (USA). The landfill is an operating mixed- municipal solid waste facility wh ere 3M sludges were disposed off. About 80 000 ng/l PFOA were de tected in the gas condensate.
Outdoor air measurements of PFO A were available from Japan a nd the USA. From the 20 samples collected around Kyoto, Japan, the highest result was 920 pg/m 3 and the lowest result was 1.6 pg/m 3 (Harada et al., 2005a). The geometric means for each of the two cities studied also varied widely: 2.0 pg/m 3 (Morioka) and 260 pg/m 3 (Oyamasaki). In the USA, samples were collected at the fence line of the Washington Wo rks manufacturing site in West Virginia (Barton et al. 2006). Of the 28 samples analyzed, 5 had levels above the reporting limit. The highest result was 900,000 pg/m 3 , and the reporting limit was between 120,000 and 170,000 pg/m 3 . No vapor-phase PFOA was detected above the detection lim it (ca. 70,000 pg/m3). These values are notably higher than the ambient levels and reporting limits from the Japanese study.

54OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Moriwaki et al. (2003) collected dust via vacuum cleaner in Japanese homes. 16 dust samples from 16 different homes were analyzed. PFOA was found inall samples at concen trations greater than the limit of determination (>50 ng/g). The PFOA con centrations in dust samples ranged from 69 - 3700 ng/g (mean: 380 ng/g, median: 165 ng/g). The sour ce for the occurrenc e of PFOA could not be identified. Archived dust samples collected by the USEPA for the CTEPP study in July 2000 to November 2001 from 56 homes in Ohio and 52 ho mes in North Carolina contained PFOA at mean concentrations of 3.155 and 2.977 μg/g dust, respectively (Strynar and Lindstrom, 2005). Shoeib et al. (2006) measured te lomer alcohols in air from the north Atlantic and Canadian Archipelago. They reported 8-2 alcohol concentrations of 5.8 – 26 pg/m 3 and 10-2 alcohol of 1.9-17 pg/m 3

. DuPont measured air concentrations using hi gh volume impact +9samplers and low volume OVS sorption tubes. The reported air concentr ations from below detection level (<0.01 ng/m 3 ) up to 75.9 ng/m 3 in ambient air samples out to two miles from their Washington Works site. (EPA-HQ-2004- 0112-0032).

Soil Limited measurements of PFOA and PFOA precursor s in soil have been reported. The most extensive data have been collected by DuPont and 3M near their facilitie s in Alabama and West Virginia. Two sets were collected near th e manufacturing sites inWest Virginia and Ohio, and the results ranged from below the reporting limit (0.17 ng/g ww) to 700 ng/g ww (DuPont, 2003c, 2005). The third set was collected at Tomakomai, a Pacific Coas t city in southern Hokkaido (Japan). As the result of an earthquake on September 26, 2003, two majo r fires occurred (26th and 28th September) at an oil storage faci lity of a refinery lo cated in the west part of Tomakomai. At least 40,000 L of fire fighting foams (FFF) that contained perfluorinated compounds were used to extinguish these fires. A monthly monitoring survey of the environmental levels of perf luoro carboxylic acids (PFCs) in the Tomakomai region was conducted between Oc tober and December 2003. Two months after the usage of FFF Yamashita et al. (2004b) detect ed 2,870 pg/g soil dw. Based on the occurrence of PFOA in snow and soil samples, the authors suggest that significant amounts of PFCs in AFFF were released into air and de posited on land by wet deposition. De Voogt and van Roon (2005) detected PFOA con centrations of 8.3 ng/g dw in soil samples collected at a contaminated site in The Netherland s (no further information provided). Oliaei et al. (2006) determined PFOA in soil samples collected under the Washingt on County landfill (USA) and in soils which were not influenced by the landfill. PFOA background concentrations of up to 4.87 ng/g were found. At the contaminated site soils showed PFOA concentrations in the range between <0.208-21.7 ng/g. Sediment and suspended matterMeasurements of PFOA in sediment samples have been reported for the Nordic countries, the Netherlands, Japan, and the USA. Schrap et al. ( 2004), and De Voogt and van Roon (2005) detected PFOA in 11 out of 33 sediment and suspended matter samples collected in the Netherlands (LOD = 0.4 ng/g dw); the single hi ghest result was 24 ng/g dw. Only2 of the 13 samples analyzed from the Nordic countries had levels above th e limit of quantitation (0.20 ng/g ww). The values were 0.28 and 0.31 ng/g ww (Kallenborn et al., 2004). Near manufacturing sites in the USA, the results were below the reporting limit of 0.2 ng/g ww (3M, 2001c), whereas results for cities with no manufacturing sites ranged from below the reporting limit to 1.8 ng/ g dw. A year later, sediment from the city reporting 1.8 ng/g dw was re-sampled, and the results were be low the reporting limit. In 2005, Oliaei et al. (2006) collected Mississippi ri ver sediment samples in the vicinity of the 3M production plant (USA). The highest PFOA concentrati on was measured at the point of discharge of

55OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

treated wastewater (18 ng/g).Sediment sample s collected under ponded water at the Washington County landfill showed PFOA concentrations of up to 31.1 ng/g. In sediment samples collected between 2002 and 2004 in the USA up to 0.625 ng/g dw we re detected (Higgins et al. 2005). In suspended sediment samples collected in the Niagara River Lucaciu et al. determined PFOA concentrations of 80 (1982), 100 (1987/88), 180 (1992/93), 220 (1995), 100 (2000), and 290 pg/g (2001). Kurunthachalam et al. (2005) and Nakata et al. (2005) report PFOA concentrations of <0.1-1.2 ng/g dw for sediment samples collected in the Kyoto area and Osaka and of 0.96 ng/g dw for sediment samples collected from the Ariake Sea (Japan). Sewage Sludge and Effluent In a sewage sludge sample collected near a manufacturing site in the USA the reported concentration of PFOA in sewage sludge was 244 ng/g dry wei ght (3M, 2001c). 2 of th e 8 samples were below the reporting limit (0.02 ng/g ww). In samples co llected away from manu facturing sites in the USA (Port St. Lucie, Florida and Cleveland, Tennessee), 3.1 ng/g dw were determined (3M, 2001c). Similarly, the highest resultfrom the Nordic countries wa s 1.1 ng/g ww (Kallenborn et al., 2004), while 8 out of 23 samples showed PFOA con centrations below 0.20 ng/ g ww. Higgins et al. (2005) analyzed a maximum PFOA c oncentration of 29.4 ng/g dw in sludge samples from the USA. PFOA was present in 11/13 samples. PFOA wasn’t detected in differe nt sludge samples collected in Germany (Schroeder 2003). Oliaei et al. (2006) and Sinclair and Kannan (2006) measured PFOA up to 241 ng/g dw in sludge samples collected at wast ewater treatment plants (WWTPs) located in St. Paul (Minnesota) and New York State. The highe st value was observed at a WWTP treating domestic, commercial, and industrial waste. In Japan the highest PFOA concentr ation in sewage effluent was m easured at the Agawam Ryuiki water disposal site (Osaka area ); 67,000 ng/l PFOA were detected. Tseng et al. (2006) detected PF OA concentrations of at maximu m 170 ng/l in indu strial and municipal WWTPs in Taiwan. Sewage effluent measurements in the USA show PFOA concentrations up to 2,400 ng/l near a manufacturing site (3M, 2001c). Sewage effluent samples co llected during 1998 to 2003 at Cottage Grove and Decatur (3M production plants; USA) showed PFOA concentrations up to 1,991,000 ng/l; measurements conducted with samples collect ed 2003 showed significantly reduced but still high PFOA concentrations (maximum 110,500 ng/l). Oliaei et al. (2006) investig ated the effluent of the 3M Cottage Grove WWTP; th ey found 62,400 ng/l. At the Washington Works Facility PFOA concentrations of max. 915,000 ng/l (2001), max. 141,000 ng/l (2002), and max. 46,400 ng/l were measured (DuPont 2003a). In the wa stewater effluent of a PTFE latex costumer of Solvay Solexis 74,000-75,100 ng/l PFOA were detected (Exygen 2003). A more recent sample from the Decatur Alabama WWTP had > 500 ppb (dw) in process sludge (3M, 2005b). Alzaga and Bayona (2004) found at maximum 4,300 ng/l PFOA in the effluent of an urbanindustrial wastewater treatment plant (WWTP) in Spain. PFOA concentrations ranging from 63-560 ng/l were found in wastewater from a printing pl ant, electronic production, leather and metal processing, paper production, a photographic laboratory, and textile processing in Austria (Hohenblum et al., 2003). PFOA levels ranging fro m 1 to 670 ng/l were detected elsewhere in the USA, Canada, and the Nordic countries. (3M 2001c; Scott et al. 2003; Berger et al. 2004; Kallenborn et al. 2004; Sinclair and Kannan 2006; Crozier et al. 2005). Remarkable results ob served Oliaei et al. (2006) who found in the effluent of a munical WWTP (S t. Paul, Minnesota, USA) higher PFOA effluent concentrations (78 ng/l) than in the influent (46 ng/l). The authors assume that precursor substances were biodegraded to PFOA during WWTP passage. Boulanger et al. (2005) analyzed in fluent, effluent, and river water at the point of discharge for a 6 million gallons per day wastewater treatment plant which receives domestic and industrial wastewa56OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

ter. This WWTP was selected because of the lack of known perfluorooctane surfactant manufacture and production. In the effluent of the WWTP, a PFOA concentration of 22 ng/l was detected (blank: PFOA concentration in HPLC grade water: 2.4 ng/l), in the influent >4 ng/l were found (exact value was not determinable due to the analytical uncerta inty encountered for this matrix-poor recovery). PFOA was measured at 8.7 ng/l in river water sample s. Boulanger et al. (2005) assume that PFOA and related compounds were introduced into WWT Ps through cleaning and care of surface-treated products (from clothing to carpets), through their use in industrial processes, and from the treatment of wastewater influent containing discarded product.Specific fluorinated analytes, perfluoroalkyl sulfonates, fluorotelomer sulfonates, perfluorocarboxylates and selected fluorinated al kyl sulfonamides were studied at ten municipal wastewater treatment plants across the US. The flow rate s in the ten plants ranged from 11,000 m 3 /day to 600,000

m 3 /day. Levels of PFOA in the influent ranged from 1.7 ng/l to 89 ng/l. The effluent concentrations of PFOA ranged from 2.5 ng/l to 97 ng/l (Schultz et.al, 2006). The effluent waters of six activated sludge wastew ater treatment plants inNew York State were monitored for PFOA and seven other perfluorinatedcompounds. All of the six plants received domestic wastewater. All but one of the plants re ceived commercial wastewater and only two plants received industrial wastewater. The levels of PFOA in the effluent of all six plants ranged from 58 ng/l to 1050 ng/l. Average concentrations of PFOA in the efflue nt of the two high capacity plants were 135 ng/l and 239 ng/l at the 25 and 35 MGD pl ants respectively. The smaller plant average effluent levels ranged from 67 ng/l to 697 ng/l. Concentrations of PF OA in the sludge samples at the two large plants were 69-241 ng/g and 18-89 ng/g (Sinclair and Kannan, 2006). Landfill Effluent From various manufacturing sites in the USA, the concentrations of PFOA in landfill effluent ranged from not detected (3M, 2001c) to 3,200,000 ng/l (DuPont, 2003a). Th e range of concentrations measured in landfill effl uent from a non-manufacturing site in the USA was 940 to 1,000 ng/l (3M, 2001c), while the range in the Nordic countries was 91 to 516 ng/l (K allenborn et al., 2004); Berger et al. 2004). Oliaei et al . (2006) detected max. PFOA c oncentrations of 81,800 ng/l in the landfill leachate sampled at the Pine Bend Landfill, Inver Grove Heights, Dakota County (USA).

Ground Water At various manufacturing sites in the USA, the levels of PFOA in ground water ranged from < LOD to 3,400,000 ng/l (Santoro 2003). In ground water potenti ally impacted by previous fire-fighting activities in the USA, the measurements ranged from below the reporting limit (which decreased from 36,000 ng/l in 1999 to 3,000 ng/l in 2003) to 6,570,000 ng/l (Moody and Field 1999; Moody et al., 2003). Ground water samples collected at Washington Co. Landfill (USA) contained PFOA concentrations ranging from 1,300-70,000 ng/l (M DH 2004). Ground water samples collected at private wells in the vicinity of the aforementioned landfill showed PF OA concentrations < 1,000 ng/l (PFOA detected but not quantified; no further information; MDH, 2004). DuPont (2003b,c) detected PFOA concentrations up to 51,000 ng/l at Chambe rs Works in Deepwater and Washington County (USA). Investigations of 3 M (2003b) revealed PFOA concentrations in ground water taken at the 3M pr oduction facilities at Decatur and Cottage Grove of max. 3,433,000 ng/l. Oliaei et al. (2006) anal yzed ground water samples collectedat the Washington County Landfill and the Pine Bend Landfill, Inver Grove Hei ghts (USA). PFOA concentrations ranged from 8- 41,600 ng/l

57OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Fresh Water PFOA in fresh water samples (e.g. la kes, rivers, streams, rain water, tap water) has been measured in the Nordic countries, Germany, The Netherlands , Japan, Taiwan, South Korea, Canada and the USA. The highest measurement from a creek afte r a spill of fire retardant foam was 11,300 ng/l, while the upstream maximum concentration wa s 33 ng/l (Moody et al., 2002). Near manufacturing facilities in the USA, the highest reported conc entration was 598 ng/l (Hansen et al., 2002), while results for some samples were below the reporting limit. In further investigations of freshwater sampled at or near 3M production facilities co ncentrations up to 10,800 ng /l were detected (3M 2005; Oliaei et al. 2006 ). Away from spills and ma nufacturing sites, the hi ghest reported concentration of PFOA in surface water was 760 ng/l, which wa s from Port St. Lucie, Florida, USA (3M, 2001c). In German rivers (Rhine, Pfinz, Neckar, Enz, Main, Kinzig, Mosel, Sieg, Erft, Ruhr, and Danube) PFOA concentrations ranging from <1 ng/l to 8 ng/l were detected (Lange et al. 2004). Skutlarek, Exner, and Färber (2006) detected <2-48 ng/l in the river Rhein (Germany; Swiss) and its most important tributaries. In the river Ruhr and its tributaries concentrations up to 3640 ng/l PFOA were detected (Skutlarek, Exner, and Färber 2006). In 2005, De Voogt and van Roon (2005) detected PFOA (160 ng/l) in freshwater samples take n from a canal after recei ving wastewater from a fire incident where fire fi ghting foam has been used. In rain water, PFOA has been detected up to 90 ng /l in a sample from Delaware, USA (Scott et al., 2006). Some rain water and surface water results we re below detection limits. In the Nordic countries PFOA concentrations up to 17 ng/l were meas ured in rain water (B erger et al., 2004; Kallenborn et al., 2004). According to Be rger et al. (2004) and Kallenbornet al. (2004) lake water sampled in the Nordic countries containe d PFOA concentrations up to 8.2 ng/l. Rainwater samples were collected in Winnipeg, Manitoba and analyzed for PFCAs. However, no PFCAs (including PFOA) were detected (method de tection limit of 0.0072 μg.L-1) (Loewen et al., 2005). Loewen et al. (2005) suggested that this may be due insufficient atmospheric concentrations of PFCAs and a relatively high method detection limit.Scott et al. (2005) determined PFOA in precipitation from 4 American and 3 Canadian locati ons; the highest value wa s 85 ng/l determined at Delaware (USA).
DuPont measured PFOA in rain at 9 locations with in two miles of their Washington Works in West Virginia in fall 2005. They detected PFOA in all samples ranging from below quantitation limit (8.8 ng/l) to 1,660 ng/l . They also m easured concentrations at a background site in Maryland at below quantitation level (11 ng/l)(EPA-HQ-2004-0112-0032). Other researchers are working on measuring rain water in North America. When their resear ch becomes available there will likely be a better understanding of the occurrence of PFOA in rain water. Saito et al. (2004) found 0.12–40 ng/l PFOA (geometric mean) in tap water samples from Osaka and the Tohoku area (Japan). Morikawa et al. ( 2005) detected 87,100 ng/l PFOA in a Japanese river. PFOA concentrations ranging from 0.9 to 62 ng/l were measured in streams of the Shihwa industrial zone and Lake Shihwa in South Kor ea (Rostkowski et al. 2005) . In Taiwanese rivers PFOA concentrations of 113-181 ng/l we re measured (Tseng et al. 2006). Scott et al. (2006b) measured PFOA and other perfluor oacids in rain from 9 sites in North America. They reported PFOA concentrations of <0.1-6.1 ng/L in remote location samples and 0.6-89 ng/L in more urban locations.

Salt Water Samples from oceans, seas and coastal waters from various regions have been analyzed for PFOA. Caliebe et al. (2004) analyzed s ea water samples collected in the German Bight and determined the concentrations and the distributi on of perfluorinated chemicals. In March 2003 at the mouth of the

58OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

river Elbe, PFOA concentrations of about 20 ng/l were detected. Along the Elbe plume towards the north, the concentration of PFOA d ecreased to ca. 6 ng/l and 4 ng/l (n ear Sylt); in the open sea the concentration dropped to ca. 1 ng/l. In the southern German Bight, PFOA concentrations of about 7 (north of Norderney) and 15 ng/l (west of Borkum) we re detected. Especially at these locations it is obvious that, in contradiction to the other measurements, the PF OA concentrations were significantly higher than the perfluorooctanesulfonate (PFOS) concentrations , indicating a different source of PFCs than at other sampling sites where PFOA and PFOS concentrations were comparable. Samples collected in July/August 2003 again showed the highest PFOA concentration (ca. 16 ng/l) at the mouth of the river Elbe. PFOA was detected at elevated concentrations near Sylt (4 ng/l PFOA) and in the southern German Bight (ca. 10 ng/l PFOA west of Borkum). Between the English and Dutch coasts (Ipswich-Rotterdam) PFOA concen trations of about 1 ng/l were found. At sampling sites in the open sea PFOA concentrations were about 0.5 ng/l. PFOA concentrations in the range between 1 and 2 ng/l were det ected at the Norwegian coast.
Within the framework of a UBA re search program, Theobald and Caliebe (2005) analyzed sea water samples collected in the German Bight and in the North Atlantic-North Polar Sea in 2004. In the region of the German Bight , the highest PFOA concentrations were detected in the river Elbe and at its mouth (Stade: 6.821 ng/l; near Cuxhaven: 3.791 ng/l). Decreasing PFOA concentrations were observed along the Elbe plume towards Sylt (Eider: 2.248 ng/l; Amrum: 2.083 ng/l; Sylt: 1.355; 1.972; and 2.182 ng/l). At one sampling station located in the direct vicinity of Sylt, the samples showed elevated PFOA concentratio ns in relation to the other two sampling poi nts near Sylt (2.182 ng/l). Elevated PFOA concentrations were also detected north of the island Schiermonnikoog (2.567 ng/l). In the open sea PFOA concentrations of < 0.2 ng/l were detected (detection limit: 0.013 ng/l). The samples collected during the cruise of the Po larstern in June/July 2004 in the North AtlanticNorth Polar Sea showed PFOA concentrations ranging from 0.035 to 0.097 ng/l (concentrations blank value corrected; The obald and Caliebe, 2005). The highest PFOA levels were observed in Asia. PFO A concentrations up to 450 ng/l were detected offshore of Japan (Saito et al., 2004). In the Atlantic Ocean even at a depth of 5,300 meters PFOA could be detected (0.05 - 0.1 ng/l; Scott et al. 2005). Background concentrations in the open oceans range from 0.06 to 0.2 ng/l (Caliebe et al., 2004; Theobald and Caliebe, 2005; Yamashita et al., 2004a; Taniyasu et al., 2004). In coastal waters from Jun-Wen estuary and Chi-Ku lagoon (Taiwan) Tseng et al. determined 130-270 ng/l PFOA.

Plants Data on PFOA uptake and occurrence in plants is very limited. Ongoing studies will provide additional information as results are finalized. Researchers at the University of Pennsylvania measured blood se rum PFOA concentrations in a stratified random sample of people residing in n ear the DuPont Washington Works facility. The reported a statistically significant correlation betw een servings of home gr own fruit and vegetable eaten and blood levels. Additionally, there was an incr ease in serum PFOA withservings of meat or game grown or harvested locally (P=0.005) ( Emme tt et al., 2006). Analysis of the concentration levels in food produce items is underway. Grass samples collected from on-site and off-site at the Washington Works DuPont West Virginia (USA) site were analyzed for PFOA. Samples we re also collected at a background location many kilometers away. They detected PFOA in all samples analyzed. On-site sampled were washed to help evaluate inthe PFOA was in the plant tissues or on the surface. On-site unwashed grass samples contai ned PFOA levels between 154 ppb and 574 ppb on a dry weight basis. Washed grass samples ha d PFOA levels between 33.5 ppb and 92.8 ppb (dry weight).

59OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

For off-site sampling DuPont collected of the a bove ground plant material and the underlying soil layer. In their report DuPont labeled these samples as 0 - 1 in. and 1 - 3 in. soil/grass samples. The 0 - 1 in. samples contained grass and the underlyi ng thatch layer. The 1-6 feet samples contained the underlying soil and root material. Off-site grass samples were collected during a dry period and following a rainfall event. Grass was sampled from 14 sites within 3.2 km of the facil ity and a background site far from the facility. Analysis of vegetation and underlying soil from all location sampled had detectable PFOA. The background samples and samples from one location n ear the facility had levels below quantitation limits (<6.0 ppb dry period and 5.0 wet period). The other 13 locations had PFOA concentrations on grass ranging 6.88 ppb - 118 ppb during the dry period and < 7.0 – 121 ppb following the rain event. The underlying soil layer material collected during the dry period between 1 -6 in. in depth contained PFOA at <7.0 to 52.8 ppb. Underlying soil fo llowing the rain even had PFOA concentrations ranging from <7 ppb to 51.6 ppb. All concen trations were on a dry weight basis (EPA-HQOPPT-0113- DuPont Quarterly MOU Status Report #2 Text, Tables and Figures). At the 3M Decatur, AL (USA) faci lity vegetation was collected in the former sludge application area, sludge application control area and background locations during October, 2004. Samples were taken of four vegetative species found in the slud ge application areas; br oomsedge, pokeweed, curly dock, and goldenrod. PFOA concentrations in the vege tation in the former industrial sludge application fields ranged from 42.3 – 912 ppb. In the sludge control and the background area that did not receive PFOA application the PFOA concentr ations ranged from 1.99 – 9.22 ppb (3M, 2005b). Samples of duckweed collected in 2005 at the site receiving the wastewater treatment discharge from the 3M Cottage Grove, MN facility c ontained PFOA at 18.10 ppb. (Oliaeie et al., 2006). In benthic algae collected at the Raisin River, the St. Clair River, and Calumet River (USA) PFOA could not be detected at a detection limit of 0.2 ng/g ww (Kannan et al., 2005). Oliaei et al. (2006) analysed 18.1 ng/g PFOA in floating Lemna minor samples collected from the river cove that receives the 3M Cottage Grove WWTP discharge.

Invertebrates Two species of invertebrates collected in the wild – mysis and diporeia – we re studied by Martin et al. (2004a), and the mean PFOA concentrations ranged from 2.5 to 90 ng/g ww. Kannan et al. (2005) analysed tissue samples of amphipods, zebra mussels, and crayfish collected at the Raisin River, the St. Clair River, and Calumet River (USA). PFOA could no t be detected in any of the samples at LODs of 0.2 - 5 ng/g ww.

Asiatic clams were collected from the same locations as freshwater fish at the 3M Decatur, AL site in December, 2004. PFOA concentrations in clam s upriver were non-quantifiable and downriver levels of PFOA ranged from 0.508 ng/g to 1.01 ng/g at Baker’s Creek (3M, 2005b). So et al. (2005a; 2006a) detect ed PFOA concentrations up to 0.660 ng/g ww in mussel and oyster tissue collected from Tokyo Bay and the east coast of China. Worm sample s (whole body) collected from Ariake Sea (western Japan) revealed PF OA concentrations of 82 ng/g ww (Nakata et al. 2005).

Freshwater Fish Over 300 samples of freshwater fish from the Nordic countries, Poland, the Czech Republic, the Netherlands, Canada and the USA have been anal yzed for PFOA (Kannan et al. 2005; Kallenborn et al. 2004; Martin et al. 2004a,b ; Schrap 2004; Sinclair et al. 2004; Giesy and Kannan 2001; Moody et al. 2002). Many of the results were below the resp ective reporting limits of these studies;

60OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

however, the relative high detectio n and quantification limits must be taken into account (LOQ up to 72 ng/g ww). The highest detected concentrati on, 91 ng/g ww, was from the liver of a fish downstream of a spill of fire retardant foam (Moody et al., 2002). More typical levels in whole samples of freshwater fish range from 1.0 to 44 ng/g wet weight (Martin et al., 2004a). PFOA was detected only occasionally (in 2 out of 30 samples) in fish liv er collected in the Nordic countries. The highest concentration is given as 1.4 ng/g ww (Kallenborn et al. 2004). Ac tual findings with improved analytical methods (reporting limits <2 ng/g ww) reveal that PFOA concentrations in fish (liver) are in the pg/l to low ng/g ww range (Oliaei et al. 2006; Sinclair et al. 2006). Plankton and Shellfish Whole samples of zooplankton, shrimp and clams fr om the Canadian eastern Arctic had maximum levels of PFOA of 3.4, 0.5 and < 0.2 ng/g ww, respectively (Tomy et al., 2004). All of the 77 oysters sampled from the Gulf of Mexico and th e Chesapeake Bay, USA, had PFOA levels below the limit of quantification (< 19 ng/g ww; Giesy and Kannan, 2001). Tsen g et al. (2006) detected 130- 160 ng/g dw and 130-80 ng/g dw in mu scle and viscous samples from oyster collected from culture sites in coastal areas of Taiwan. Nakata et al . (2005) found PFOA concentr ations up to 9.5 ng/g ww in soft tissue of oyster, mussel, and clam coll ected at tidal flat from Ariake Sea (Japan).

Saltwater Fish
PFOA has been measured in hundreds of saltwater fi sh from the Baltic Sea, the Mediterranean Sea, the North Pacific Ocean, from the coasts of the Canadian Arctic, Taiwan, Japan, Belgium, the Netherlands, Nordic Countries; Poland, Colombia and Antarctica. Oliver o-Verbel et al. (2005) determined PFOA concentrations up to 370 ng/ml in bile. In fish sample s collected in the Ariake Sea (Japan) liver tissue contained PFOA concentrati on of at maximum 15 ng/g ww (Nakata et al. (2005). Liver and muscle samples of coastal fish co llected in Taiwan reveal ed PFOA concentrations of 120-340 ng/g dw and 100 ng/g dw, respectively (Tse ng et al. 2006). The highest level detected in a liver sample from a fish from a Nordic count ry was 5.4 ng/g ww (Kalle nborn et al., 2004), while the highest level detected in a liv er sample from a fish from th e Canadian Arctic was 5.3 ng/g ww (Tomy et al., 2004). PFOA concentrations up to 46 ng/ g ww were determined in muscle tissue from fish sampled in Belgium (Giesy and Kannan, 2001). Reptiles and amphibians Wild turtles that rank high in th e food chain that were captured in a river in Japan in 2003 contained serum levels of PFOA up to 115.6 ug/l and were found to bioconcentrate the PFOA in the surface waters in which they li ved by an average (geometric mean) fact or of 3.2, with a ra nge of BCFs from 0.8 to 15.8. Serum concentrations of PFOA from 94 turtles ranged from 0.3 to 115.6 ug/l. Surface water concentrations in the ar ea where the turtles lived ranged from .021 to 87.1 ug/l of PFOA. (Morikawa et.al, 2005). Plasma samples from sea turtles along the south eastern coast of the USA were analyzed for PFOA (Keller et al., 2004; 2005). The mean results were between 3 and 4 ng/ml. Kannan et al. (2005) investigated tissue samples of turtles and frogs colle cted in the USA. At LODs of 2.5 (turtle) and 72 ng/g ww (frog) PFOA could not be detected. Morikawa et al. ( 2005) analyzed serum samples of turtles caught in a Japanese river. PFOA concentrations ranging from <0.2 to 870 ng/ml were detected.

Birds The highest PFOA concentration was detected in a liver sample from a cormorant from the Mediterranean Sea (450 ng/g ww; Corsolini and Kannan, 2004; Kannan et al., 2002b). It should be noted though that this value appears to qu alify as an outlier as the concentration was 4.5 times greater than

61OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

the standard deviation of the mean. In liver sa mples from glaucous gull, black-headed gull, and black-eared kite PFOA was found occasionally in maximum concentrations of 21 ng/g ww (Tomy et al., 2004; Kannan et al., 2002a; Berger and Ha ukas 2005; Nakata et al. 2005). Double-crested cormorant (blood and yolk samples) and sea eagle (blood samples) showed PFOA concentrations up to 245 ng/ml (Giesy and Kannan, 2001). In spleen of brown pelicans co llected from Colombia PFOA concentrations of 182 ng/g ww were found (Olivero-Verbel et al. 2005). Hölmström et al. (2005) analysed guillemot egg samples collected between 1968 and 2003 in the Ba ltic Sea; PFOA could not be dete cted at a reporting limit of 3 ng/g ww. Osprey nesting areas in the Balt imore Harbor, the Patapsco River, the Anacostia and the middle Potomac Rivers, the Elizabeth River of Maryland and Virginia (USA) and what was thought to be a reference site were monitored for a suite of cont aminants included perfluorinated chemicals in 2000 and 2001. The C9, C10, C11, and C12 perfluorocarboxcy lic acids were detected in all egg samples monitored in these areas but only quantifiable in a few (values were <30 ng/g wet weight). PFOA was not detected in any samples (LOD = 30 ng/g) (Rattner et al. 2004).

Mammals Over 1,000 samples of terrestrial (e.g., mice, mink) and marine (e.g., seals, whales, dolphins) mammals have been analysed for PFOA. The highest result for liver samples from mink was 108 ng/g ww (Giesy and Kannan, 2001), and the highest result for liver samples from otters was 19 ng/g ww (Hoff et al., 2004). For seals, the highest result for a liver sample was 41 ng/g ww (Giesy and Kannan, 2001). The highest mean result for live r samples from polar bears, 55.8 ng/g ww, was from South Baffin Island (Smithwick et al., 2005a). Smithwick et al. 2006 analyzed archived polar bearliver tissue samples from two geographic locations in the North American Arctic collected from 1972-2002. The eastern group (northern Baffin Island, Canada) comprised 31 samples, the wester n group (Barrow, Alaska) 27 samples. Between 1972-2002, in the eastern group a significant increase in PFOA concentration was observed (1972: ca. 3 ng/g ww PFOA; 2002: ca. 20 ng/g ww PFOA). On ly slight changes in PFOA concentrations were detected in liver tissue samples collected at the western location (1972: ca. 2 ng/g ww PFOA; 2002: ca. 5 ng/g ww PFOA). The highest level of PFOA measur ed in liver samples from dol phin, 20 ng/g wet weight, was from the Gulf of Mexico, and the high est blood level obtained for sample s collected on the east coast of USA was 163 ng/g ww (Houde et al. 2005) and 4 ng/ml for dolphin from th e Mediterranean Sea (Giesy and Kannan, 2001). Van de V ijver et al. (2004) analysed 62 ng/g ww PFOA in the tissue of harbor porpoises sampled as by-caught from coas tal waters from Iceland, Norway, Denmark, and the German Baltic Sea. In liver samples of whale PFOA was only occasionally detected at concentrations in the pg/g to low ng/g ww range (Bossi et al. 2005; Corsolini and Kannan 2004; Kallenborn et al. 2004; Tomy et al. 2004; van de Vijver et al. 2003; Kannan et al. 2002b). Guruge et al. (2005a) analysed serum and liver samples of domestic animals (cattle, ch icken, pig, and horse). In serum samples up to 965 pg/ml (chicken) while in liver sample s up to 300 pg/g (pig) were detected. De Silva and Mabury (2004) isolated isomers of pe rfluorocarboxylates in pol ar bears from two geographical locations, the southeastern Hudson Ba y region of Canada (2002) and central eastern Greenland (1999 - 2001). According to the study au thors hypothesis branched PFOA isomers were found in all seven Greenland bear samples. Conversely, the PFOA inall eight Canadian bears consisted solely of the linear isomer. These contrast ing patterns indicate a difference in the source of

62OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

PFOA. A complete absence of branched PFOA in the Canadian bears sugge sts a non-ECF input of PFOA to this location. In the Greenland polar bear samples, the average branched PFOA isomer composition was 5.0% of the total PFOA. This value is mu ch less than the 22% found in the ECF standard. Volatile telomer compounds may be responsible for the additional linear PFOA determined in the samples as PFOA and perfluorononanoic acid (PFNA) are the major products of the photochemically induced degradation of 8:2 fluorotelomer alcohol . The authors suggest that the di fference of perfluorocarboxylic acids (PFCs) occurrence is due to the air current. Whereas the impact in North America is due to releases there, Greenland can be influenced from both North American and European sources. This may account for the differences in the consistent and distinct pattern of PFCA isomer distribution between Canadian and Greenland polar bear samples. This implies that there are differences in production and/or application of fluorinated materials between these two regions.

The environmental distribution a nd the biomagnification of a suit e of perfluoroalkyl compounds (PFCs), was investigated in th e food web of the bottlenose dolphi n. Dolphin tissue samples and surficial seawater were collected at Sarasota Bay, Florida and Charleston Harbor, South Carolina. Dolphin plasma PFOA concentrations for Sarasota averaged 3.4 ng/l (for discussion of biomagnification and bioaccumulation see text above “ 2.2.6 Bioaccumulation”) (Houde et al., 2006). PFOS and PFOA levels in serum of the giant pandaand the red panda were taken in animals captured in zoos and animals parks from six provinces in China. The PFOA level ranged from 0.33 to 8.20 ug/l for the red panda and from 0.32 to 1.56 ug/l for the giant panda. Greater concentrations of the fluorochemicals were found in panda sera for those individuals collected from zoos near urbanized or industrialized areas thanother areas (Dai et al., 2006). Prevedouros et al. (2006) suggested that patterns in chain lengths and the branching of perfluorinated substances in environmental samples may pr ovide clues to their sour ces. For instance, they note that branched materials ar e only known to arise from the EC F production processes. However, associating chain length patterns inenvironmental samples to patterns in sources is complicated because patterns may become altered as a result of di fferent partitioning and uptake/clearance rates in biota for the various perfluorinat ed substances. Clearance rates of linear and branched PFOA in rats are different, with increased clearance of the branched material as compared to lin ear (Loveless et al, 2006). Hence, occurrence of only linear PFOA in bi ota does not seem to provide definitive information on sources. The environmental levels of PFOA are compiled in Table 20. As this is a hazard assessment and not a risk assessment, this table contains a list of published data of PFOA emissions found in the environment without interpretation to relevance.

Other Media The US EPA collected archived dust samples for PFOA in July 2000 to November 2001. The samples were obtained from 56 homes in Ohio and 52 homes in North Carolina. The samples from the 56 homes in Ohio contained PFOA at a mean concentration of 3.155 μg/g. Samples from the 52 homes in north Carolina contained PFOA at a meanconcentration of 2.977 μg/g (Strynar and Lindstrom 2005). Archived EPA house dust samples(n=112) were siev ed to 150 um and analyzed for select PFCs. Telomer alcohols (6:2 , 8:2, 10:2 FTOH) were an alyzed by GC/MS after sonic extraction & SPE cleanup. PerF carboxlic ac ids(C6-12) and sulfonates (PFOS, PFHS, PFBS) were analyzed by LC/MS/MS. Mean concentrations ranged from 0.357ug/g (PFDA) to 11.23 ug/g (PFHS). Maximum ranged from 2.42 ug/g (6:2 FTOH) to 357.0 ug/g (PFHS) ( Strynar and Lindstrom, 2005).

63OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Drinking water from surface water: Billot (2005) detected PFOA in drinking water samples (Oakdale Municipal Well (USA) in concentrations of <0.2-858 ng/l. Skutlarek, Exne r, and Färber (2006)investigated drinking water sampled in the Ruhr area in Germany; PFOA was present in the samples in concentrations ranging fr om <0.2 to 519 ng/l. The water s upplies are mainly based on bank filtration from surface water (see above page 58 ). PFOA was measured in polar ice caps from three areas in the high Arctic (Melville Ice Cap, Northwest Territories; Agassi z Ice Cap, Nunavut; and De von Ice Cap, Nunavut) (Young et al ., 2005). PFOA concentrations ranged from 0.00245 to 0.00315 μg.L -1 suggesting that contamination may be a result of atmospheric input. However, it shoul d be noted that these values were not blankcorrected due to the unavailability of blanks at th e time of analysis. Therefore, actual results may be lower than reported. A flux calculation was performed using a total area of the Canadian Arctic and sub-Arctic as 3.38 E6 km 2 and precipitation and snow density da ta from the Melville Ice Cap. A concentration of 0.001 μg.L -1 PFOA was assumed in the Melville Ice Cap samples. The authors based this concentration on the assumption that the estimated flux falls in th e range of atmospheric chemistry models based on the oxidation chemistry of FTOH alcohols in the atmosphere. Thesemodels estimated the total flux of PFCAs to the Arctic on the order of 0.1 to 1 tonne/year. The resulting atmospheric flux was 450 kg/year, which is comparable to values calculated from atmospheric modeling

64OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Outdoor AirEPA-HQOPPT-0113- 0032 Site No data 0.09 ng/tube OVS samples limit of detection (0.09- 0.64 ng/tube fraction) to 60.2 ng/m 3 HVS samples 0.01 to 75.9 ng/m 3 2005; Air samples were collected with OSHA Versatile Sampling (OVS) method tubes and High-Volume Sampling (HVS) equipment within a two mile radius ofthe Washington Works, DuPont site in West Virginia; USA Oliaei et al., 2006 Site NR NR About 80,000 ng/l 2005; gas condensate collected from the condensation of water vapor in the gas; condensates sampled at the Pine Bend Landfill, Inver Grove Heights, Dakota County, Minnesota, USA; the landfill is an operating, mixed-municipal solid waste facility where 3M sludges were disposed off Barton et al., 2006 Site 5/28 120,000 – 170,000 pg/m 3 < RL – 900,000 pg/m 3 2003 – 2004; fence line monitoring at Washington Works, Parkersburg, West Virginia, USA
Amb 8/8 0.46 pg/m 3 geom. mean = 2.0 pg/m 3 (1.6 – 2.6 pg/m 3 ) Morioka, Japan Harada et al., 2005a

Amb 12/12 0.33 pg/m 3 geom. mean = 260 pg/m 3 (72 – 920 pg/m 3 ) Oyamasaki, Japan, whichhas a major truck route passing through; corres-ponding concentration in airborne dust geom. mean = 3,400 ng/g
Indoor Air
Moriwaki et al., 2003 Site 16/16 50 ng/g 69 - 3,700 ng/g dust(mean = 380 ng/g; median 165 ng/g Vacuum cleaner dust samples from 16 different homes in Japan
Strynar and Lindstrom, 2005 Amb No data/108 No data mean concentrations
3.155 and 2.977 μg/g dust 200-2001; archived dust samples collected at 56 homes in Ohio and 52 homes in North Carolina, USA

65OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

SoilEPA-HQOPPT-0113-DuPont Quarterly MOU Status Report #2 Text, Tables and Figures Site No data No data Soil samples collectedfrom the 3”-6” intervals 72-130 ng/g (ppb)
soil samples in the 6”- 12” interval 44-130 ng/g (ppb) Soil samples were collectedon-site at the DuPont Washington Works West Virginia site (USA) from soil 3”-6”and 6”-12” below the surface

Site NR 0.208 ng/g <0.208-21.7 ng/g Soil samples collected under the Washington County landfill at different depths (0-26 ft), USA Oliaei et al. 2006

Amb NR 0.208 ng/g 1.28-4.87 ng/g Background soils were samples from an area upgradient of the groundwater direction of flow and out of the direct influence of the area wheregroundwater was spray irrigated at the Washington County landfill (depths: 0-8 ft), USA De Voogt and van Roon, 2005 Site NR 0.4 ng/g dw 8.3 ng/g dw 2004-2005 ; soil samplescollected from a contaminated site ; The Netherlands
Site No data 0.2 ng/g (ppb) 29-3140 ng/g (ppb; 0- 3”) 34.2-5690 ng/g (ppb; 3- 6”) 18.5-1870 ng/g (ppb; 6- 12”) 25.3-8780 ng/g (ppb; 1- 10”) 14.1-3800 ng/g (ppb; >10”) 2006; soil samples collectedat depth of 0-12”’ from former sludge application area, formerly investigated areas and background areas of the 3M Decatur facility, AL; USA

Amb No data 0.2 ng/g (ppb) 0.2-9.92 ng/g (ppb; background level) 3M, 2005b

Site No data 0.2 ng/g p(pb) 0.2-810 ng/g (ppb; formerly investigated areas) DuPont, 2005b Site 9/9 2.1 – 21 ng/g dw 2.2 – 700 ng/g dw 2003; 0 – 1 foot deep; West Virginia, USA
Yamashita et al., 2004b Site 2/2 NR mean = 2780 pg/g dw 2003; two months after firefighting foam used at refinery fire; Tomakomai Bay, Japan DuPont, 2003c Site 11/22 0.17 – 0.19 ng/g

ww < RL – 170 ng/g ww Two borings up to 56 feet deep; Washington County, Ohio, USA

66OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Sediment 3M, 2005b Site No data/12 No data 0.393 and 24.1 ng/g 2006; sediment samples collected in the Tennessee River adjacent to the 3M Decatur, AL facility; USA Oliaei et al., 2006 Site NR <0.298-0.301 ng/g 18 ng/g (sedimentcove; point of discharge of treated wastewater) 6.62 ng/g (immediately downstream of the cove)<0.301-1.31 ng/g (concentration decresedwith increasing distance from the discharge location1) <0.298 ng/g (upstream of point of discharge) 2005; Mississippi river sediment samples collected at a depth of 10 cm in thevicinity of the 3M production plant;USA

Oliaei et al. 2006 Site NR NR 22,3-31.1 ng/g Sediment samples collected under ponded water at the Washington County landfill, USA De Voogt and van Roon, 2005 Site NR 0.4 ng/g dw 7.0 ng/g dw 2004-2005 ; sediment samples collected from a contaminated site ; the Netherlands Higgins et al., 2005 Amb 15/17 0.011 ng/g dw RL – 0.625 ng/g dw 2002-2005 sediments collected from San Francisco Bay area, Palo Alto MudFlats, Hayward, CA, Baltimore, MD, and Corvallis, OR, USA
Kurunthachalam et al., 2005 Amb NR 0.1 ng/g dw <0.1-1.2 ng/g dw 2003-2005; sediment samples collected from rivers in the Kyoto area and from Osaka; Japan Lucaciu et al., 2005 Amb NR NR Ca. 80 pg/g (1982) Ca. 100 pg/g (1987/88) Ca. 180 pg/g (1992/3) Ca. 220 pg/g (1995) Ca. 100 pg/g (2000) Ca. 290 pg/g (2001) 1981-2001; suspended sediment samples collected at Niagara-on-the-Lake in the Niagara River

Nakata et al., 2005 Amb 5/5 0.09-0.10 ng/g dw 0.96 ng/g dw Sediment samples collectedfrom the Ariake Sea, western Japan

67OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Sediment (continued)
Kallenborn et al., 2004 Amb 2/13 0.20 ng/g ww < RL – 0.31 ng/g ww 2003/2004; sediment samples collected in industrialized or urban areas of Finland, Sweden, Norway, Denmark, Iceland, Faeroe Islands Schrap et al., 2004; De Voogt and van Roon, 2005 Amb 11/33 0.4 ng/g dw < RL – 24 ng/g dw 2004; sediment and suspended matter from fresh water and marine locations; the Netherlands

Amb 6/6 0.2 ng/g ww 0.30 – 1.8 ng/g dw 1999; Port St. Lucie, Florida, USA
Amb 0/6 0.2 ng/g ww < RL 2000; Port St. Lucie, Florida, USA
Amb 0/6 0.2 ng/g ww < RL Cleveland, Tennessee, USA 3M, 2001c

Site 0/24 0.2 ng/g ww < RL Manufacturing sites (Decatur, Alabama; Mobile, Alabama; Columbus, Georgia; Pensacola Florida), USA Sewage Sludge Oliaei et al., 2006 Site NR NR 3.79 ng/g dw (primary sludge)21.5 ng/g dw (secondary sludge)11.1 ng/g dw (dewatered sludge prior to incineration) 2005; sludge samples collected at metro WWTP a municipal wastewater treatment located in St. Paul, Minnesota; USA

Sinclair and Kannan, 2006 Site 10/10 10 ng/g dw 69-241 ng/g dw (D,C, I) 18-89 ng/g dw (D,C) 2004-2005; sludge samples collected from two WWTPs located in New York State, USA; WWTP treating: domestic (D), commercial (C) and industrial waste (I) Higgins et al., 2005 Amb 11/13 1 ng/g dw < RL – 29.4 ng/g dw 1998-2004; Pacific NW, West, West North Central, West South Central, and Northeast; USA
Kallenborn et al., 2004 Amb 15/23 0.20 ng/g ww < RL – 1.1 ng/g ww 2003/2004; sewage samples collected in industrialized or urban areas of Finland, Sweden, Norway, Denmark, Iceland, Faeroe Islands

68OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Sewage Sludge (continued) Schroeder, 2003 Site NR 10 mg/kg dw < RL sludge samples collected in NRW, Germany
Amb 2/6 0.2 ng/g ww < RL – 3.1 ng/g dw Port St. Lucie, Florida, and Cleveland, Tennessee; USA 3M, 2001c

Site 6/8 0.2 ng/g ww < RL – 244 ng/ g dw Manufacturing sites (Decatur, Alabama; Mobile, Alabama;Columbus, Georgia; Pensacola Florida), USA Sewage Effluent Oliaei et al., 2006 Site NR NR 78 ng/l (effluent) 46 ng/l (influent) 2005; influent and effluent of metro WWTP a municipal wastewater treatment located in St. Paul, Minnesota; USA Oliaei et al., 2006 Site NR NR 62,400 ng/l 2005; effluent of 3M CottageGrove WWTP treating process wastewater generated from the production facility; after passing three differenttreatment systems the combined effluents are discharged to the Mississippi River; USA Schultz et.al, 2006 Site No data No data Influent 1.7-89 ng/l effluent 2.5-97 ng/l Sewage effluent at ten municipal wastewater treatment plants across the USA

Sinclair and Kannan 2006 Site 45/45 2.5 ng/l 142-398 ng/l (D,C, I) 66-202 ng/l (D,C) 435-851 ng/l (D,C) 361-1050 ng/l (D,C) 132-196ng/l (D,I) 58-78 ng/l (D,C) 2004-2005; WWTP effluent samples collected from six plants located in New York State, USA; WWTP treating: domestic(D), commercial © and industrial waste

Tseng et al. 2006 site 2/2 0.8 ng/l Not detected-39 ng/l Industrial effluents of textile and dyeing manufacturers in Chung-Li, Taiwan Tseng et al. 2006 site 2/2 0.8 ng/l 36-170 ng/l effluents of municipal WWTPs near Taipei and Tainan, Taiwan
Site NR/3 < 2 ng/l Mean: 22 ng/l 2004; WWTP effluent in Iowa City; USA Boulanger et al., 2005
Site NR/3 < 2 ng/l Mean: 8.7 ng/l 2004; river water concentration at point of discharge; Iowa City, USA
Crozier et al. 2005 Site NR 1 ng/l

0.1 ng/g dw 10-34 ng/l (stp effluent)0.7-13 ng/g dw (biosolids) Sewage treatment plant final effluent and biosolids

69OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Sewage Effluent (continued)
Site NR/15 100 ng/l Mean: < 100 – 1,400 ng/l (range: 100 – 4,300 ng/l) 2003; effluent of an urbanindustrial WWTP; Spain Alzaga and Bayona, 2004

Site 1/5 100 ng/l 150 ng/l 2003; 5 locations in Barcelona harbor; Spain Berger et al., 2004 Amb 5/5 NR Median: 20.5 ng/l samples collected in industrialized or urban areas of Finland, Sweden, Norway, Denmark, Iceland, Faeroe Islands
Site 2/2 No data Influent 26,000 and 194,000 ng/l 2003; Influent concentrations of PFOA to the Chambers Works wastewater treatment plant; USA
Site 2/2 No data Effluent 279,000 and 475,000 ng/l, 2003; effluent concentrations of PFOA directly from the Chambers Works wastewater treatment plant; USA EPA, 2004b – AR226-2620

Site No data No data 102,000 and 193,000 ng/l (April 18/ May 16) 76,000, 144,000 and 97,000 ng/l (June 10/11) 2003; PFOA discharged fromthe Chambers Works wastewater treatment plant (USA)following the addition of noncontact cooling water to the Delaware river
Kallenborn et al., 2004 Amb 7/7 0.06 ng/l 1.3 – 23 ng/l 2003/2004; samples collectedin industrialized or urban areas of Finland, Sweden, Norway, Denmark, Iceland, Faroe Islands Saito et al., 2004 Site NR 0.1 ng/l 67,000 ng/l 2003; sewage effluent at Aigawa Ryuiki water disposal site; Osaka area, Japan 3M, 2003b; Santoro 2003 AR226-1482 Site NR NR 1,991,000 ng/l (Jan.- March 2000) 216,000 ng/l (Sept.- Oct. 2000) 180,000 ng/l (Dec, 2002) 17,700-110,500 ng/l (average values for Jan.-June 2003) 2000-2003; effluent wastewater samples collected at theplant outfall on the Mississippi River; 3M productionplant Cottage Grove, Minnesota, USA

70OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Sewage Effluent (continued) 3M, 2003b Site NR 0.082 ng/l 602,000 ng/l (1998) 766,000 ng/l (1999) 1,028,000 ng/l (2000) 310,000 ng/l (2001) 58,000-88,300 ng/l (Jan./May 2003) all values refer to average concentrations 1998-2003; effluent wastewater samples collected at Decatur, Alabama, USA

Site 14/14 50 ng/l 750-46,400 ng/l (2003)
Site 83/84 50 ng/l <50-141,000 ng/l (2002) DuPont 2003a

Site 26/26 50 ng/l 118-915,000 ng/l (2001) 2001-2003; six outfalls were sampled at Washington Works Facility; USA

Exygen 2003 Site 2/2 50 ng/l 74,000-75,100 ng/l 2003; wastewater of a PTFE latex costumer of Solvay Solexis; USA
Hohenblum et al., 2003 Site 4/15 50 ng/l 63 – 560 ng/l wastewater from printing plant, electronic production, leather processing, metal processing, paper production, photo-graphic laboratory, textile production; Austria Scott et al., 2003 Amb 7/7 NR 3 – 30 ng/l 3 Great Lakes

Amb 6/6 25 ng/l 40 – 670 ng/l Port St. Lucie, Florida, and Cleveland, Tennessee; USA 3M, 2001c

Site 8/8 25 ng/l 84 – 2,400 ng/ l Manufacturing sites (Decatur, Alabama; Mobile, Alabama;Columbus, Georgia; Pensacola Florida), USA
Landfill Effluent Oliaei et al., 2006 Site NR NR 61,650 ng/l (average; range: 14,200-81,800 ng/l) 2005; Landfill leachate sampled at the Pine Bend Landfill, Inver Grove Heights, Dakota County, Minnesota,USA; the landfill is an operating, mixed-municipal solid waste facility where 3M sludges were disposed off Berger et al., 2004 Amb 6/6 NR Median: 297 ng/l samples collected in industrialized or urban areas of Finland, Sweden, Norway, Denmark, Iceland, Faeroe Islands

71OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Landfill Effluent (continued)
Kallenborn et al., 2004 Amb 9/9 0.06 ng/l 91 – 516 ng/l 2003/2004; samples collectedin industrialized or urban areas of Finland, Sweden, Norway, Denmark, Iceland, Faeroe Islands DuPont, 2003a Site 168/168 50 ng/l 61 – 3,200,000 ng/l 1994–2003 monitoring; Parkersburg, West Virginia, USA
Amb 4/4 25 ng/l 940 – 1,000 ng/l Port St. Lucie, Florida, USA 3M, 2001c
Site 3/8 25 ng/l < RL – 48,000 ng/l Manufacturing sites (Decatur, Alabama; Mobile, Alabama;Columbus, Georgia; Pensacola Florida), USA
Ground Water 3M, 2005b Amb/ Site No data 25 ng/l 2270-3,857,000 ng/l (sludge application area) 220–94,700 ng/l (control sludge application wells) 25 (non-detect) - 1,333,000 ng/l (LOI PFOA levels) 2006; groundwater samples collected from the former sludge application area, the sludge application control andthe area wells at the 3M Decatur, AL facility; USA

EPA-HQOPPT-0113- 0032). Site No data/70 No data <0.050 to 3.22 x 10 5 μg/l (APFO 1996- 2006) 537-118,000 ng/l (PFOA 2004-2006 1996-2006; groundwater wells (on the plant site) hasbeen monitored at the Washington Works facility in West Virginia (USA)EPA-HQOPPT-0113- 0032). Site No data 50 ng/l (APFO) PFOA and APFO levels at the Local landfill 141-69,000 ng/l (PFOA) <50-91,100 ng/l (APFO) Dry Run landfill <50-32,600 ng/l (PFOA) <50-39,000 ng/l (APFO) PFOA and APFO levels at the Letart landfill 126-20,200,000 ng/l (PFOA) 58-57,200,000 ng/l (APFO) 1996-2006; Groundwater monitoring wells located at three different landfill sites in the vicinity of the DuPont, West Virginia Washington Works plant were monitored

Oliaei et al., 2006 Site NR NR 41,600 ng/l Ground water at the Wahington County Landfill, USA

72OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Ground Water (continued) Oliaei et al., 2006 Site NR NR 8-1,600 ng/l 2005; ground water from monitoring wells sampled at the Pine Bend Landfill, Inver Grove Heights, Dakota County, Minnesota, USA; the landfill is an operating, mixed-municipal solid waste facility where 3M sludges were disposed offSkutlarek, Exner, and Färber, 2006 Amb NR 2 ng/l <0.2-519 ng/l Ca. 2005; drinking water samples collected in the Ruhr area, Germany Billot, 2005 Amb 10/12 25 ng/l <25-858 ng/l (average values) 2004; drinking water samplescollected at Oakdale Municipal Well; USA Billot, 2005 Amb 8/8 25 ng/l 288-856 ng/l (average values) 2005; drinking water samplescollected at Oakdale Municipal Well; USA
Site No data 50 ng/l 191 to 5230 ng/l 2003; groundwater from 6 monitoring wells located on the Chambers Works site EPA, 2004b – AR226-2620

Site No data 50 ng/l 354 to 46,800 ng/l 2003; Groundwater from 5 interceptor wells
Site NR NR 1,300 – 70,000 ng/l 2003/2004; Ground watersamples collected at Washington Co. landfill, USA MDH, 2004

Amb 7/32 NR < 1,000 ng/l (PFOAdetected but not quantified) 2003/2004; ground waterfrom private wells in the vicinity of the Washington Co. landfill, USA 3M, 2003b Site NR 82 ng/l 87-3,433,000 ng/l (2001; average) <LOD-2,386,000 ng/l (2003; average) 2001/2003; groundwater well samples collected at Decatur, Alabama, USA

73OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Ground Water (continued)
Amb NR NR 174,000 ng/l (2002; eastern/downgradientof site industrial activities)
Amb NR NR 4,630-6,410 ng/l(2001;western/ downgradient of site industrial activities)
Site NR NR 5,290-6,330 ng/l (2001; central to site industrial activities 3M, 2003b

Site NR NR 1,185-40,100 (2001; site-wide ground water from the production wells) 2001-2002; groundwater samples collected at and near 3M production plant (Cottage Grove, Minnesota), USA

Amb NR NR 314 ng/l (upgradient of site industrial activities)
Site NR NR 10,200 ng/l (central to site industrial activities
Amb NR NR 4,810 ng/l (western/ downgradient of site industrial activities)
Amb NR NR 136,000 ng/l (eastern/downgradient of site industrial activities) 3M, 2003b

Site NR NR 27,900 ng/l (site-wide ground water from the production wells) 2003; groundwater samplescollected at and near 3M production plant (Cottage Grove, Minnesota), USA

DuPont, 2003b Site 14/14 10 ng/l 200 – 47,000 ng/l Chambers Works in Deepwater, New Jersey, USA DuPont, 2003c Site 85/87 10 ng/l < RL – 51,000 ng/l 2001–2003; Washington County, Ohio, USA Moody et al., 2003 Site 9/10 3,000 ng/l < RL – 105,000 ng/l 5 years after fire-fighting activity; Wurtsmith Air Force Base, Michigan, USA
Site 14/14 NR 310 – 174,000 ng/l 2001–2003; Cottage Grove, Minnesota, USA Santoro, 2003

Site 14/16 NR < RL – 3,400,000 ng/l 2001, 2003 ; Decatur, Alabama, US

74OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Ground Water (continued) DuPont, 2002 Site 43/62 10-50 ng/l < RL – 8,700 ng/l
max. concentation in drinking water: 6,500 ng/l max, concentration in unused water: 8,700 ng/l 2002; well, cistern, and spring water samples collected within a two-mile radius ofthe Washington Works facility; Washington County, Ohio, USA

Site 8/14 36,000 ng/l means = < RL – 6,570,000 ng/l Naval Air Station Fallon, Nevada, USA Moody and Field, 1999
Site 9/11 36,000 ng/l means = < RL – 116,000 ng/l Tyndall Air Force Base, Florida, USA
Fresh Water 3M, 2005b Site No data/12 No data 73.4 to 420 ng/l 2006; surface water samples collected in the Tennessee River and a nearby creek called Baker’s Creek adjacentto the 3M Decatur, AL facility; USA 3M, 2005b Site No data/16

No dat/30 4.92 ng/l 4.92-26 ng/l (longitudinal)

2.66-435 ng/l (transect samples) 2006; surface water samples(collected in longitudinal direction along the length of the river, n=16; collected across the river, n=30) collected in the Tennessee River and a nearby creek called Baker’sCreek adjacent to the 3M Decatur, AL facility; USAEPA-HQOPPT-0113- 0032 DuPont Quarterly MOU Status Report #2 Text, Tables and Figures Amb No data 1 ng/l public supply 1-71,900 ng/l

residential well samples 5-27,100 ng/l 2002/2004; DuPont has been sampling public water supply systems in West Virginia (n=12) and Ohio (n=7) for APFO and PFOA since 2004. Additionally, DuPont hassampled twenty-five residential water wells in the vicinity of Washington Works on a quarterly basis for APFO and PFOAEPA-HQOPPT-0113- 0032 DuPont Quarterly MOU Status Report #2 Text, Tables and Figures EPA-HQ-2004- 0112-0032 Site 11/11 8.8 ng/l <8.8-1,660 ng/l
(seven samples were non-quantifiable, four ranged between 10.2 and 52.9 ng/l and one measured 1660 ng/l) 2005; precipitation samples were collected at the DuPont Washington Works site, West Virginia, USA

75OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Fresh Water (continued) EPA-HQ-2004- 0112-0032 Amb No data 11 ng/l <11 ng/l 2005; DuPont measured PFOA in rain at a background site in Maryland; USA

Oliaei et al., 2006 Site NR 4.9 ng/l 3,250 ng/l (at 3M discharge location; cove water)3,650 ng/l (at 3M discharge location; cove water top surface) 35.3 ng/l (immediately downstream of the 3M discharge)<4.9 ng/l (other downstream locations) 2005; Mississippi river samples collected 2 ft below the river surface in the vicinity of the 3M production plant in Minnesota, USA

Sinclair et al., 2006 Amb 51/51 2.5 ng/l 10-173 ng/l 2004; water samples collected from nine major water bodies of New York State, USASkutlarek, Exner, and Färber, 2006 Amb NR 2 ng/l <2 – 48 ng/l Ca. 2005; surface water samples collected in the river Rhine (Schaffhausen, CH tillWesel, Niederrhein, Germany) and its most important tributaries (N eckar, Main, Nahe, Lahn, Mosel, Ahr, Sieg, Wupper, Erft, Ruhr, Emscher, and Lippe: sample collection at the mouth of the rivers)Skutlarek, Exner, and Färber, 2006 Amb NR 2 ng/l <0.2-3640 ng/l Ca. 2005; surface water samples collected in the riverRuhr and its tributaries, Germany Amb 3/3 0.8 ng/l 113 ng/l (mean) Water samples from TourChyan river, Taiwan Tseng et al., 2006
Amb 1/1 0.8 ng/l 181 ng/l Water samples from Nan-Kan river, Taiwan 3M, 2005a Site 3/5
3/5 Ca. 25 ng/l <25-10,800 ng/l <25-9,100 ng/l 2005; surface water sampled within the 3M Guin (USA) facility at two locations De Voogt and van Roon, 2005 Site 2/2 80 ng/l 160 ng/l 2004-2005; freshwater samples taken from a canal after receiving wastewater from afire incident where fire fighting foam has been used; the Netherlands
Furdui et al. 2005 Amb 30/33 1-4 ng/l 2-7 ng/l 2004; water samples collected from the Great Lakes (11 locations)

76OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Fresh Water (continued) Kannan et al., 2005 Amb NR NR 4 – 14.7 ng/l 2001; water samples from Raisin River and St. Clair River, Michigan, USA
Kurunthachalam et al., 2005 Amb NR 7.9 ng/l 7.9-110 ng/l 2005; river water samples collected in the Kyoto area ; Japan Loewen et al., 2005 Amb NR 7.2 ng/l. <7.2 ng/l. 2004; rain water samples collected in Winnipeg; Canada
Morikawa et al., 2005 Amb NR 0.1 ng/l 16.7-87,100 ng /l 2003; surface water samples from a Japanese river
Rostkowski et al., 2005 Site 27/27 0.9 ng/l 0.9-62 ng/l 2004; water samples were collected from streams of the Shihwa industrial zone and Lake Shihwa, South Korea Scott et al., 2005 Amb NR NR Max. 85 ng/l (Delaware) 1998-1999 (American locations); precipitation samples from 4 American sites (New York, Delaware, Maryland,and Vermont-urbanized locations) and 3 Canadian sites (Saturna Island, Algoma and Kejimkujik-rural locations)
Simcik and Dorweiler, 2005 Amb 36/36 0.29-0.58 ng/l 0.14-0.66 ng/l (remote lakes) 0.45-19 ng/l (urban surface waters) 0.28-3.4 ng/l (Lake Michigan) Surface water samples collected in the mid-western US (Minneapolis, remote areas of Minnesota, shore of LakeSuperior, and Voyageurs National Park along the US_Canadian boder)
Amb 4/4 NR median: 7.8 ng/l lake water samples collectedin industrialized or urban areas of Finland, Sweden, Norway, Denmark, Iceland, Faeroe Islands Berger et al., 2004

Amb 5/5 NR 13.1 ng/l rain water, samples collectedin industrialized or urban areas of Finland, Sweden, Norway, Denmark, Iceland, Faeroe Islands
Amb 8/8 13 ng/l 21 – 47 ng/l Lake Erie, Canada/USA Boulanger et al., 2004 Amb 8/8 13 ng/l 15 – 70 ng/l Lake Ontario, Canada/USA

77OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Fresh Water (continued)
Amb 6/6 0.06 ng/l 8.2 – 17 ng/l 2003/2004; rain water samples collected in industrialized or urban areas of Finland, Sweden, Norway, Denmark, Iceland, Faeroe Islands Kallenborn et al., 2004

Amb 5/5 0.06 ng/l 4.8 – 8.2 ng/l Lake water; samples collectedin industrialized or urban areas of Finland, Sweden, Norway, Denmark, Iceland, Faeroe Islands EPA, 2004b – AR226-2620 Amb No data 10 ng/l <10-577 ng/l 2003; surface water from Delaware river Lange et al., 2004 Amb NR 1 ng/l < RL – 8 ng/l German rivers Rhine, Pfinz, Neckar, Enz, Main, Kinzig, Mosel, Sieg, Erft, Ruhr and Danube
Amb 79/79 0.1 ng/l 0.1 – 460 ng/l Rivers in 6 districts throughout Japan Saito et al., 2004
Amb NR/30 0.1 ng/l geom. mean: 0.12 - 40 ng/l 2003; tap water; Osaka and Tohoku area, Japan
Amb NR/44 8 ng/l < RL – 36 ng/l 2001-2003; 9 locations throughout Michigan, USA Sinclair et al., 2004
Amb NR 8 ng/l < RL – 16 ng/l Background
Site 2/2 NR mean = 160 ng/l 2003; one month after fire fighting foam used at refinery fire; runoff into Tomakomai Bay, Japan Yamashita et al., 2004b

Site 4/4 NR mean = 18 ng/l 2003; two months after fire fighting foam used at refinery fire; snow near Tomakomai Bay, Japan DuPont, 2003b Site NR NR max. = 89 ng/l Deepwater, New Jersey, USA
Amb NR/43 0.5 ng/l < RL – 90 ng/l Rain water archived from 1998; New York, Maryland and Delaware, USA
Amb NR/16 0.5 ng/l < RL – 4 ng/l Rain water from 2002; Algoma, Canada
Amb 14/14 0.5 ng/l 1 – 10 ng/l 2001; tributaries to the Great Lakes Scott et al., 2003

Amb NR 0.5 ng/l < RL – 1 ng/l Mid depth; 4 Great Lakes

78OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Fresh Water (continued)
Amb 0/20 25 ng/l < RL Upstream from fluorochemical manufacturing facility; Tennessee River, Decatur, Alabama, USA Hansen et al., 2002

Site 18/20 25 ng/l < RL – 598 ng/l Downstream from fluorochemical manufacturing facility; Tennessee River, Decatur, Alabama, USA
Site 48/51 9 ng/l < RL – 11,300 ng/l 2000/2001; spill of fire retardant foam; Etobicoke Creek, Toronto, Canada Moody et al., 2002

Amb 8/8 9 ng/l mean = 20 ng/l (8 – 33 ng/l) 2000/2001; upstream of spill

Amb 4/4 25 ng/l 97 – 760 ng/l Port St. Lucie, Florida, USA 3M, 2001c
Site 3/8 25 ng/l < RL – 63 ng /l Manufacturing sites (Decatur, Alabama; Mobile, Alabama;Columbus, Georgia; Pensacola Florida), USA
Salt Water Houde et al., 2006 Amb No data No data 3.6 ng/l (Sarasota sea; averaged) 9.5 ng/l (Charleston; averaged) Sea water samples collected from the Sarasota Bay and Charleston; USA

Tseng et al., 2006 Amb 2/2 0.8 ng/l 130-270 ng/l Coastal waters collected from Jun-Wen estuary and Chi-Ku lagoon, Taiwan
site 2/2 ca. 3 ng/l 62.4-63.7 ng/l 2003; 3 months after firefighting foam was used; Tomakomai Bay (Japan) Taniyasu et al., 2005

site 2/2 ca. 3 ng/l 34.8-35.1 ng/l 2004; Tokyo Bay (Japan) Scott et al., 2005 Amb NR NR 0.050 – 0.10 ng/l Atlantic Ocean; depths to 5,300 meters
Amb 14/14 0.013 ng/l 0.140 – 6.82 ng/l 2004; Elbe River estuary and German Bight Theobald and Caliebe, 2005
Amb 12/12 0.008 ng/l 0.035 – 0.097 ng/l 2004; North Atlantic, North Polar Sea Nakata et al., 2005 Amb 5/5 0.8-10 ng/l 50 ng/l water samples collected from the Ariake Sea, western Japan Berger et al., 2004 Amb 11/11 NR median: 5.2 ng/l 2003/2004; water samples collected in industrialized orurban areas of Finland, Sweden, Norway, Denmark, Iceland, Faeroe Islands Caliebe et al., 2004 Amb 18/19 0.5 ng/l approx. 0.5 – 20 ng/l 2003; Elbe estuary and German Bight (North Sea)

79OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Salt Water (continued)
Kallenborn et al., 2004 Amb 15/15 0.06 ng/l 3.5 – 8.5 ng/l 2003/2004; sea water samples collected near industrialized or urban areas of Finland, Sweden, Norway, Denmark, Iceland, Faeroe Islands Saito et al., 2004 Amb 6/6 0.1 ng/l 1.9 – 450 ng/l 2003; Japan (Sea of Kushiro, Mutsu Bay, Honjo Marina, Funabashi Coast, Pacific Ocean, Koshien Coast) So et al., 2004 Amb 31/31 0.02 ng/l 0.24 – 320 ng/l 2003/2004; coastal waters of Hong Kong, South China and Korea
Amb 14/14 0.013 ng/l 0.140 – 6.82 ng/l 2004; Elbe River estuary and German Bight Theobald and Caliebe, 2005
Amb 12/12 0.008 ng/l 0.035 – 0.097 ng/l 2004; North Atlantic, North Polar Sea
Amb 3/3 approx. 0.01 ng/l 154 – 192 ng/l 2002; Tokyo Bay; depth 0 – 2 meters
Amb NR/4 approx. 0.01 ng/l 0.137 – 1.060 ng/l Offshore of Japan
Amb NR/12 approx. 0.01 ng/l 0.673 – 5.450 ng/l Coastal area of Hong Kong
Amb NR/14 approx. 0.01 ng/l 0.243 – 15.300 ng/l Coastal area of China
Amb NR/10 approx. 0.01 ng/l 0.239 – 11.300 ng/l Coastal area of Korea
Amb 3/3 approx. 0.01 ng/l 0.10 – 0.15 ng/l 2002; Mid Atlantic Ocean; depth 0 – 2 meters
Amb NR/9 approx. 0.01 ng/l 0.100 – 0.439 ng/l Mid Atlantic Ocean Yamashita et al., 2004a; Taniyasu et al., 2004
Yamashita et al., 2005

Amb 5/5 approx. 0.01 ng/l 0.076 – 0.51 ng/l 2002; Sulu Sea; depth 10 – 3,000 meters
Amb 2/2 approx. 0.01 ng/l 0.16 – 0.42 ng/l 2002; South China Sea; depth 10 meters
Amb 2/2 approx. 0.01 ng/l 0.136 – 0.142 ng/l 2002; Western Pacific Ocean; depth 0 – 2 meters
Amb 5/5 approx. 0.01 ng/l 0.015 – 0.056 ng/l Central to Eastern Pacific Ocean; depth 0 – 4,400 meters Yamashita et al., 2004a Taniyasu et al., 2004
Yamashita et al., 2005

Amb NR/9 approx. 0.01 ng/l 0.160 – 0.338 ng/l North Atlantic Ocean
Site NR/6 NR mean = 4.15 ng/l 2003; one month after fire fighting foam used at refinery fire; Tomakomai Bay, Japan Yamashita et al., 2004b

Site NR/8 NR mean = 2.67 ng/l 2003; two months after fire fighting foam used at refinery fire; Tomakomai Bay, Japan

80OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Plants
site No data No data 42.3–912 ng/g (former sludge application area) 3M, 2005b

Amb No data No data 1.99–9.22 ng/g (sludge application control areaand background locations) 2004; 3M Decatur, AL facility collected vegetation in its former sludge applicationarea, sludge application control area and background locations (plants sampled: broomsedge, pokeweed, curly duck, goldenrod)
Site No data 6 ng/g dw (ppb; dry period) 5.0 ng/g dw (ppb; wet period) On-site unwashed grass samples 154-574 ng/g dw Washed grass samples 33.5-92.8 ppb dw.EPA-HQOPPT-0113-DuPont Quarterly MOU Status Report#2 Text, Tables and Figures Amb No data 6 ng/g dw (ppb; dry period) 5.0 ng/g dw (ppb; wet period) background samples and samples from onelocation near the facility<6.0 ng/g dw (dry period)<5.0 ng/g dw (wet period) PFOA concentrations at the other 13 locations

-on grass 6.88-118 ng/g dw (dry period) < 7.0 – 121 ng/g dw (following the rain event)
-underlying soil layermaterial collected during the dry period <7.0 - 52.8 ng/g dw (1”-6” in depth)
-Underlying soil following the rain event <7 - 51.6 ng/g dw

all concentrations were on a dry weight basis Grass samples collected fromon-site and off-site at Washington Works DuPont West Virginia site, USA
Off-site grass samples were collected during a dry period and following a rainfall event; grass was sampled from 14 sites within 3.2 km of the facility and a background site far from the facility

Oliaei et al., 2006 Site NR NR 18.1 ng/g 2005; floating Lemna minor samples collected from the river cove that receives the3M Cottage Grove wastewater treatment plant discharge, Minnesota; USA Kannan et al., 2005 Amb NR 0.2 ng/g ww < RL 1999; benthic algae sampled at the Raisin River, the St. Clair River and Calumet River; USA

81OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Invertebrates 3M, 2005b Site No data No data Upriver non-quantifiable downriver 0.508-1.01 ng/g 2004; asiatic clams were collected from locations (Baker’s Creek) at the 3M Decatur plant, AL; USA So et al., 2005a; So et al., 2006a Amb NR 0.204 ng/g ww <204-328 ng/g ww 2004; mussel tissue samples collected from six locations along the east coast of China

So et al., 2005a; So et al., 2006a Amb NR 0.204 ng/g ww 660 ng/g ww 2004; oyster tissue samples collected from Tokyo Bay, Japan

Kannan et al., 2005 Amb NR 0.2/5 ng/g ww < RL 1998/1999; tissue samples of amphipods, zebra mussels, and crayfish collected at the Raisin River, the St. Clair River and Calumet River; USA Nakata et al., 2005 Amb 5/5 3 ng/g ww 82 ng/g ww Worm samples (whole body) collected from Ariake Sea, western Japan Martin et al., 2004a Amb NR 2 ng/g ww means = 2.5 – 90 ng/g

ww 2001; whole body homogenates of Mysis and diporeia; Lake Ontario, Canada/USA
Freshwater Fish 3M, 2005b Site No data No data Upriver non-detect - 5.88 ng/g (filet and whole body) Downriver non-quantifiable - 6.06 ng/g (filet and whole body) 2004; two species of freshwater fish, channel catfish and large mouth bass were collected from the TennesseeRiver and Creeks in the vicinity of the 3M Decatur plant, AL; USA
Site No data No data PFOA levels in viscerasamples for both species 1.23-1440 ng/gPFOA in carcass samples 0.412-10.8 ng/g PFOA in filet samples 0.234-9.17 ng/gEPA-HQOPPT-0113- 0032 DuPont Quarterly MOU Status Report #2 Text, Tables and Figures

Amb No data No data PFOA in background fish -viscera 0.85-42.6 ng/g

carcass 0.087-1.59 ng/g filet 0.146-1.65 ng/g Channel catfish and large mouth bass were collected from the Ohio River near the Washington Works, DuPont WVA site (USA)

82OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Freshwater Fish (continued) Oliaei et al., 2006 Site 13/20 0.345-0.374 ng/g

ww <0.345-1.27 ng/g ww 2004; fish liver samples (small mouth bass, white bass, common carp, walleye) collected in the Mississippiriver in the immediate vicinity of the 3M Cottage Grove discharge; USA Oliaei et al., 2006 Site 0/20 0.417-0.668 ng/g

ww <0.417-<0.668 ng/g

ww 2004; fish fillet samples (small mouth bass, white bass, common carp, walleye) collected in the Mississippiriver in the immediate vicinity of the 3M Cottage Grove discharge; USA Oliaei et al., 2006 Site NR/>100 0.236-0.944 ng/ml <0.236-15.50 ng/ml 2005; fish blood samples (small mouth bass, white bass, common carp, walleye, small mouth buffalo, gizzardshad, emerald shiner) collected in the Mississippi river in the immediate vicinity ofthe 3M Cottage Grove discharge and in the Lake Pepin area; USA Sinclair et al., 2006 Amb ca. 60/66 1.5 ng/g ww <1.5-7.7 ng/g ww 2001-2003; liver samplesfrom smallmouth and largemouth bass collected in New York State, USA
Amb NR 0.2/2 ng/g ww < RL 1998/1999; tissue samples of round gobies and smallmouth bass collected at the Raisin River, the St. Clair River and Calumet River; USA Kannan et al., 2005

Amb NR/26 18/72 ng/g ww < RL Tissue samples of various fish (salmon, trout, carp, Lake whitefish) from Michigan waters, USA
Kallenborn et al., 2004 Amb 2/30 1 ng/g ww < RL – 1.4 ng/g ww 2003/2004; various species; liver samples collected near industrialized or urban areas of Finland, Sweden, Norway, Denmark, Iceland, Faeroe Islands Martin et al., 2004a Amb NR 2 ng/g ww means = 1.0 – 44 ng/g

ww 2001; alewife, rainbow smelt, slimy sculpin and lake trout; whole homo-genate samples; Lake Ontario, Canada/USA Martin et al., 2004b Amb 0/5 2.0 ng/g < RL Trout and lake whitefish;liver samples; Canadian Arctic

83OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Freshwater Fish (continued) Schrap, 2004 Amb 0/NR 3 ng/g ww < RL Eel filet samples (fresh water); the Netherlands Sinclair et al., 2004 Amb 0/87 72 ng/g ww < RL 1999/2000; liver and muscle samples from various fish; Michigan, USA
Amb 1/7 8.0 ng/g ww < RL – 13.7 ng/g ww Various species; egg samples; Michigan, USA
Amb 3/103 8.0 ng/g ww < RL – detected (not quantified) Various species; muscle samples; Michigan, USA Giesy and Kannan, 2001; as sum-marized in APME, 2002

Amb 0/159 72 ng/g ww < RL Various species; whole, liver, muscle, egg, blood and testes samples; Michigan, USA, Poland and Czech Republic
Site 8/8 1.2 ng/g ww 6.0 – 91 ng/g ww 2000/2001; after spill of fire retardant foam; commonshiner; liver samples; Etobicoke Creek, Toronto, Canada Moody et al., 2002

Amb 1/1 1.2 ng/gww 88 ng/g ww 2000/2001; upstream of spill;common shiner; liver samples; Etobicoke Creek, Toronto, Canada
Plankton and Shellfish
Amb 3/3

3/3 10 ng/g dw 130 ng/g dw (mean)

160 ng/g dw (mean) Muscle samples from oysters (Crassostrea gigas) collected from oyster culture sites in Chi-Ku coastal area; Taiwan Tseng et al., 2006

Amb 3/3

3/3 10 ng/g dw 130 ng/g dw (mean)

180 ng/g dw (mean) Viscus samples from oysters (Crassostrea gigas) collected from oyster culture sites in Chi-Ku coastal area; Taiwan Nakata et el., 2005 Amb 16/16 3 ng/g ww 6.0-9.5 ng/g ww Soft tissue of oyster, mussel, and clam collected at tidal flatfrom Ariake Sea, western Japan Amb 5/5 0.2 ng/g ww mean = 2.6 ng/g ww (1.7 – 3.4 ng/g ww) 2002; zooplankton; whole samples; Canadian eastern Arctic
Amb 3/7 0.2 ng/g ww mean = 0.17 ng/g ww (< RL–0.5 ng/g ww) 2000/2001; shrimp; whole samples; Canadian eastern Arctic Tomy et al., 2004

Amb 0/5 0.2 ng/g ww < RL 2002; clams; whole samples; Canadian eastern Arctic Giesy and Kannan, 2001; as sum-marized in APME, 2002 Amb 0/77 19 ng/g ww < RL American oyster; whole samples; Gulf of Mexico and Chesapeake Bay, USA

84OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Saltwater Fish Bossi et al., 2005 Amb 0/20 12 ng/g ww < RL Shorthorn sculpin; liver samples; Greenland
Amb 6/6 10 ng/g dw 120 ng/g dw (tilapia; mean) 340 ng/g dw (seaperch; mean) Liver samples of coastal fish (tilapia; Japanese seaperch); Taiwan Tseng et al., 2006

Amb 3/3 10 ng/g dw 100 ng/g dw (mean) Muscle samples of coastal fish (tilapia); Taiwan
Gulkowska et al., 2005 Amb NR/18 0.05 ng/ml <0.05-0.73 ng/ml 2002-2003; cod samples (blood) collected from the Gulf of Gdansk, Baltic Sea (Poland)
Kurunthachalam et al., 2005 Amb NR 2.5 ng/g ww <2.5-10 ng/g ww 2005 ; fish (scad, jack mackerel, sand fish, trout, sardine)were purchased in retail shopping markets and liver samples investigated; Japan Nakata et al., 2005 Amb NR/21 3 ng/g ww <3-15 ng/g ww Liver samples of fish (filefish, sea bream, red sea bream,right eye flunder, hammerhead sharkcollected from shallow water at Ariake Sea, western Japan
Amb -/10 3 ng/g ww <3 ng/g ww Liver of herbivore mudskipper collected from tidal flat at Ariake Sea, western Japan Nakata et al., 2005

Amb 3 ng/g ww 7.8 ng/g ww Liver of omniivore mudskipper collected from tidal flat at Ariake Sea, western Japan
Olivero-Verbel et al., 2005 Amb NR/47 50 ng/ml <50-370 ng/ml (bile) 2003; mullet (mugil incilis) were collected along the northcoast of Colombia (3 locations) and tissue and organs investigated
Corsolini and Kannan, 2004 Amb 0/15 72 ng/g ww < RL Northern blue fin tuna and swordfish; liver and blood samples; Mediterranean Sea Johnson et al., 2004 Amb NR NR a Mullet; bile samples; Cartagena Bay, Colombia
Kallenborn et al., 2004 Amb 2/14 1 ng/g ww < RL – 5.4 ng/g ww Various species; liver samples; samples collected near industrialized or urban areas of Finland, Sweden, Norway, Denmark, Iceland, Faeroe Islands Martin et al., 2004b Amb 0/5 2.0 ng/g < RL Arctic sculpin, northern pikeand white sucker; liver samples; Canadian Arctic

85OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Saltwater Fish (continued) Schrap, 2004 Amb 0/NR 3 ng/g ww < RL Flounder and cod filet samples; the Netherlands
Amb 1/6 0.2 ng/g ww mean = 0.16 ng/g ww (< RL – 0.5 ng/g ww) 2000/2001; arctic cod; whole samples; Canadian eastern Arctic Tomy et al., 2004

Amb 2/7 0.2 ng/g ww mean = 1.2 ng/g ww (< RL – 5.3 ng/g ww) 2000/2001; deepwater redfish; liver samples; Canadian eastern Arctic Kannan et al., 2002b Amb 0/35 72 ng/g ww < RL 1999, swordfish, tuna, Atlantic salmon; liver samples; Baltic and Mediterranean Seas
Amb 8/26 7.5 ng/g ww < RL – 46 ng/g ww Various species; muscle samples; Belgium Giesy and Kannan, 2001; as sum-marized in APME, 2002 Amb 0/65 72 ng/g ww < RL Various species; whole, liver and blood samples; North Pacific Ocean, Baltic Sea,Mediterranean Sea and Antarctica Reptiles and amphibians
Amb NR/5 2.5 ng/g ww < RL Snapping turtle tissue samples collected from Macomb County, Michigan, USA Kannan et al., 2005

Amb NR/4 72 ng/g ww < RL Green frog tissue samples collected from Kalamazoo, Michigan, USA
Morikawa et al., 2005 Amb 91/94 0.2 ng/ml <0.2-870 ng/ml 2003-2004; serum samples of turtles caught in a Japanese river
Amb 73/73 NR 0.49-8.14 ng/ml (mean = 3.2 ng/ml) Juvenile loggerhead sea turtles; plasma samples; southeastern coast USA Keller et al., 2004; Keller et al., 2005
Amb 6/6 NR 2.77-4.25 ng/ml (mean = 3.6 ng/ml) Juvenile Kemp’s ridley sea turtles; plasma samples; south-eastern coast USA
Birds Berger and Haukas 2005 Amb NR 0.128 ng/g ww <0.128 ng/g ww 2004; glaucous gull liver samples collected from themarginal ice zone in the Barents Sea; Norway Bossi et al., 2005 Amb 0/38 12 ng/g ww < RL Black guillemot and fulmar; liver samples; Greenland and Faeroe Islands
Holmström et al., 2005 Amb 0/146 3 ng/g ww < RL Guillemot; egg samples from 1968–2003; Baltic Sea

86OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Birds Kannan et al., 2005 Amb 0/NR 19/38 ng/g ww < RL 2000; Bald eagle samples of liver, kidney, gall bladder, muscle, testes, ovary; Upper Peninsula, Michigan, USA Nakata et al., 2005 Amb -/13 3 ng/g ww <3 ng/g ww Liver of mallard and blackheaded gull sampled from shallow water at Ariake Sea, western Japan
Olivero-Verbel et al., 2005 Amb 1/5 30 ng/g ww 182 ng/g ww (spleen) 2004; brown pelicans were collected in Cartagena Bay (Colombia) and tissue and organs investigated
Corsolini and Kannan, 2004 Amb 12/12 2.5 ng/g ww 29 – 450 ng/g ww 1997; common cormorant; liver samples; Mediterranean Sea
Kallenborn et al., 2004 Amb 0/2 1 ng/g ww < RL 2002; fulmar; egg samples collected near industrialized or urban areas of Finland, Sweden, Norway, Denmark, Iceland, Faeroe Islands Martin et al., 2004b Amb 0/15 2.0 ng/g ww < RL Common loon, northern fulmar and black guillemot; liver samples; Canadian Arctic
Amb 0/4 0.2 ng/g ww < RL Black-legged kittiwake; liver samples; Canadian eastern Arctic Tomy et al., 2004

Amb 1/5 0.2 ng/g ww mean = 0.14 ng/g ww (< RL – 0.3 ng/g ww) Glaucous gulls; liver samples; Canadian eastern Arctic Kannan et al., 2002a Amb 4–8/83 19 ng/g ww (Japan)36 ng/g ww (Korea) < RL – 21 ng/g ww Black-headed gull, blackeared kite; liver samples; Japan and Korea

Kannan et al., 2002b Amb 12/12 1–72 ng/g ww 29 – 450 ng/g ww Cormorant; liver samples; Baltic and Mediterranean Seas

87OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Birds (continued)
Amb 15/18 19 ng/g ww < RL – 444 ng/g ww Various fish-eating waterbirds; liver samples; Mediterranean Sea, Midway Atoll and Japan
Amb 3/88 30 ng/ml < RL – 49 ng/ml Double crested cormorant and sea eagle; blood samples;Great Lakes, Japan and Mediterranean Sea
Amb 2/4 180 ng/ml < RL – 245 ng/ml Double crested cormorant; yolk samples; Lake Winnipeg Giesy and Kannan, 2001; as sum-marized in APME, 2002

Amb 0/349 37 ng/g ww < RL Various fish-eating water birds; liver, kidney, muscle, ovary, testes, gall bladder, blood, serum, plasma and egg samples; USA, Poland, BalticSea, Mediterranean Sea, Japan and India
Mammals 3M, 2005b Site 2/15

0/15 400 ng/ml

- <1000 ng/ml serum) non-detect (liver) 2004; small mammal Hispid cotton rat: livers and serum)sampling was conducted concurrently with vegetation sampling at the 3M Decatur, AL facility; USA Dai et al., 2006 Amb No data No data 330-8200 ng/l (red panda) 320-1560 ng/l (giant panda) PFOA levels in serum of the giant panda and the red pandawere taken in animals captured in zoos and animals parks from six provinces in China

Smithwick et al., 2006 Amb 58/58 2.3 ng/g ww Eastern group: ca. 3 ng/g ww (1972) ca. 20 ng/g ww 2002) Western group: ca. 2 ng/g ww (1972) ca. 5 ng/g ww (2002) 1972-2002; polar bear liver tissue samples collected from two geographic locations in the North American Arctic(eastern group: northern Baffin Island, Canada; western group: Barrow, Alaska Guruge et al., 2005a Amb 49 pg/ml sera 49 pg/g liver Serum: 50 pg/ml (cattle) <49-965 pg/ml (chicken) 176 pg/ml (pig) <49 pg/ml (horse) liver: <49-239 pg/g (cattle) <49 pg/g (chicken) <49-300 pg/g (pig) Serum and liver samples of domestic animals (cattle,chicken, pig, and horse); Japan

Houde et al., 2006 Amb No data No data Dolphin plasma3.4 ng/l (Sarasota; averaged) 43 ng/l (Charleston; averaged) Dolphin tissue samples were collected at Sarasota Bay,Florida and Charleston Harbor, South Carolina; USA

88OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Mammals (continued)

Amb 13/13 0.5 ng/g ww 0.7-26 ng/g ww Amb 2/2 0.5 ng/g ww 0.6-0.9 ng/g ww
Amb 42/42 0.5 ng/g ww 1-70 ng/g ww
Amb 47/47 0.5 ng/g ww 4.6-163 ng/g ww Houde et al., 2005

Amb 5/5 0.5 ng/g ww 20-115 ng/g ww 2003, blood samples from bottlenose dolphin collected in the Gulf of Mexico and the Atlantic Ocean along the coast of the USA
-Sarasota Bay -Bermuda -Indian River Lagoon -Charleston -Delaware Bay
Nakata et al., 2005 Amb 5/5 3 ng/g ww 9.1 ng/g ww Liver of finless porpoise sampled from shallow water at Ariake Sea, western Japan
Smithwick et al., 2005a Amb > 75% 2.3 ng/g ww mean = 2.4 – 36 ng/g

ww min = < 2.3 – 20 ng/g

ww max = 9.04 – 55.8 ng/g

ww 2002; polar bear samples collected from 5 locations in North American Arctic and 2 locations in European Arctic

Taniyasu et al.,2005 ; Taniyasu et al., 2005a Amb 4/4 0.01-0.08 ng/g ww 0.28-0.29 ng/g ww Beaver liver ; Poland

van de Vijver et al., 2005 Amb -/82 62 ng/g ww <62 ng/g ww 2002; harbor seals stranded along the coastline of the Wadden Sea were collectedand organs and tissues analysed (liver, kidney, blubber, muscle, tracheo-branchial muscle, and spleen); The Netherlans
Amb 0/10 12 ng/g ww < RL Polar bear; liver samples; Greenland Bossi et al., 2005
Amb 0/35 12 ng/g ww < RL Mink whale and long-finned pilot whale; liver samples; Greenland and Faeroe Islands Kannan et al., 2005 Amb NR/8 2 ng/g ww 6.3 – 40 ng/g ww Mink liver samples; Kalamazoo, Michigan, USA
Amb NR/29 2 ng/g ww mean = 10 ng/g ww 1999-2001; polar bear; liver samples; Greenland Smithwick et al., 2005b
Amb 0/30 12 ng/g ww < RL Ringed seal; liver samples; Greenland
Corsolini and Kannan, 2004 Amb <4/19 2.5 ng/g ww < RL – 3.8 ng/g ww Dolphin and whale; liver, muscle and blood samples; Mediterranean Sea

89OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Mammals (continued) Ellis et al., 2004 Amb NR NR 10 – 110 ng/g ww (value refers to PFOAand/or PFNA; not further specified) Polar bear; liver

De Silva and Mabury 2004 Amb NR/15 NR Greenland>Canada 1999-2002; polar bear; liversamples; Greenland and Canada Guruge et al., 2004 Amb 12/15 0.010 ng/ml < RL – 0.10 ng/ml Black beef cattle followed from age 9 to 27 months; plasma samples; Japan Hoff et al., 2004 Amb 0/42 110 ng/g ww < RL Wood mice; liver samples; Belgium
Amb 8/8 1 ng/g ww 0.3 – 5.6 ng/g ww Seal; liver; samples collected near industrialized or urban areas of Finland, Sweden, Norway, Denmark, Iceland, Faeroe Islands
Amb 0/5 1 ng/g ww < RL Mink whale; liver; samples collected near industrialized or urban areas of Finland, Sweden, Norway, Denmark, Iceland, Faeroe Islands Kallenborn et al., 2004

Amb 4/4 1 ng/g ww 0.35 – 1.7 ng/g ww Pilot whale; liver; samples collected near industrialized or urban areas of Finland, Sweden, Norway, Denmark, Iceland, Faeroe Islands
Amb 0/39 2.0 ng/g ww < RL Mink, fox and ringed seal;liver samples; Canadian Arctic Martin et al., 2004b

Amb 7/7 2.0 ng/g ww mean = 8.6 ng/g ww (2.9 – 13 ng/g ww) Polar bear; liver samples; Canadian Arctic

Amb NR/5 0.2 ng/g ww mean = 0.34 ng/g ww (< RL – 0.7 ng/g ww) 1998; walrus; liver samples; Canadian eastern Arctic
Amb 5/5 0.2 ng/g ww mean = 0.9 ng/g ww (0.7 – 1.1 ng/g ww) 2000; narwhal whale; liver samples; Canadian eastern Arctic Tomy et al., 2004

Amb 5/5 0.2 ng/g ww mean = 1.6 ng/g ww (1.0 – 2.8 ng/g ww) 1996; Beluga whale; liver samples; Canadian eastern Arctic Van de Vijver et al., 2004 Amb 1/41 62 ng/g ww 62 ng/g ww Harbor porpoises liver tissue; sampled as by-caught fromcoastal waters; Iceland, Norway, Denmark, German Baltic Sea

90OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 20 Environmental levels of PFOA (continued)

Reference Site or Amb. Detection Frequency Reporting Limit PFOA Concentration Comments

Mammals (continued) Van de Vijver et al., 2003 Amb 0/184 110 ng/g ww < RL 1995–2000; porpoise, seal, dolphin and whale; kidney and liver samples; stranded along Dutch, Belgian and French North Sea coasts
Amb <10/52 19 ng/g ww < RL–39 ng/g ww Seal; liver samples; Baltic Sea
Amb 0/15 72 ng/g ww < RL 1991–1999; dolphin andwhale; liver and muscle samples; Baltic and Mediterranean Seas Kannan et al., 2002b

Amb <4/4 2.5 ng/ml mean = 3.1 ng/ml (< RL – 3.8 ng/ml) 1997; bottlenose dolphin; blood samples; Mediterranean Sea
Amb 23/112 4.5 ng/g ww < RL – 27 ng/g ww 1995–2000; mink; liver samples; Illinois, Massachusetts,South Carolina and Louisiana, USA Kannan et al., 2002c

Amb <11/20 7.5 ng/g ww < RL – 19 ng/g ww 1997/1998; river otter; liver samples; Washington and Oregon, USA
Amb 18/31 19 ng/g dw < RL – 108 ng/g dw Mink; liver samples; Massachusetts, USA
Amb 3/87 19 ng/g ww < RL – 41 ng/g ww California sea lion and ringedseal; liver samples; California, USA, and Baltic Sea Giesy and Kannan, 2001; as summarized in APME, 2002

Amb 11/62 5.0 ng/ml < RL – 23 ng/ml Seal; blood and serum samples; Baltic, Caspian and Mediterranean Seas
Amb 0/51 36 ng/g ww < RL Mink and river otter; liver samples; 5 states in USA
Amb 0/215 72 ng/g ww < RL Various marine species; liver,brain, kidney and blood samples; USA, Canada and Antarctica
Amb 1/3 7.5 ng/g ww < RL – 20 ng/g ww Dolphin; liver samples; Gulf of Mexico
Amb 3/4 2 ng/ml < RL – 4 ng/ml Dolphin; blood samples; Mediterranean Sea Giesy and Kannan, 2001; as summarized in APME, 2002

Amb 0/31 38 ng/g ww < RL Dolphin and whale; liver andmuscle samples; India, Mediterranean Sea, Gulf of Mexico and Atlantic Ocean a = concentration reported as “several hundred ng/ml” Amb = ambient environmental samples (not collected near a site of high contamination) APME = Association of Plastic Manufacturers in Europe dw = dry weight

91OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

geom. = geometric LOQ = limit of quantitation NR = not reported RL = reporting limit, which is the detection limit in some st udies and the limit of quantitation in other studies (the value for the reporting limit is included in parenthesis) Site = samples collected near a manufacturing site or a highly contaminated site (e.g., where fire fighting foam was previously used) ww = wet weight

2.3 Human Exposure

2.3.1 Human Biomonitoring Data PFOA has been measured in human serum in relatively small numbers of samples from the general population in several countries th roughout the world and in serum of workers occupationally exposed to APFO. Most of the biomonitoring data on PFOA have been collected in the U.S. and in Japan. In general, report ed mean serum PFOA concentrations in the general U.S. population have been lower than those reported in U.S. workers e xposed to PFOA. However, data recently submitted to US EPA have reported subs tantially higher PFOA concentrations in U.S. residents living near a facility that uses PFOA than those that have been reported thus far in the general population. Small numbers of blood samples have also beencollected, and PFOA has been detected, in presumably non-occupationally exposed subjects in Ca nada, Columbia, Poland, Belgium, India, Korea, Sri Lanka, Japan, and Sweden. Mean concentrations inthese countries are generally similar to those in the U.S., with the exception of Korea where PFOA concentrations were much higher. However, it is currently difficult to comp are or interpret PFOA concentra tions in human blood across the globe without nationally representative sa mples and standard analytical methods. The mechanisms and pathways leading to the pr esence of PFOA in human blood ar e not well charac terized, although PFOA has been detected in wildlife, air, water, soil, indoor dust, sludge, and food.

2.3.2 Occupational Exposure 3M and DuPont have been the primary U.S. produ cers and users of perfl uorinated compounds. Both companies offer voluntary medical surveillance progr ams to workers at plants that produce or use perfluorinated compounds. 3M disconti nued manufacturing PFOA from 2000 to 2002.

2.3.2.1 3M Occupational Data 3M has been offering voluntary medical surveillan ce to workers at plants that produce or use perfluorinated compounds since 1976. Serum PFOA levels in 3M workers have been measured since 1993 (Olsen et al., 2003a; 2003b; 2003c; 2001a; 2001b; 2001c; 2000; 1999; 1998c). Prior to this time, analytical capabilities precluded the accura te measurement of any specific fluorochemical analyte and only total organic fluorine was measur ed. PFOA analysis differed slightly in each surveillance program year and different laboratories were used to assay PFOA in each year. The samples were analyzed for PFOA using high perfor mance liquid chromatography mass spectrometry (HPLC/MS), but the extraction methods differed slightly each year (O lsen, et al., 2001f; 2003a, 2003b). Serum PFOA concentrations for workers particip ating in 3M’s biomonitoring program have been reported for 3 plants: Cottage Grove, Minnesota ; Decatur, Alabama; and Antwerp, Belgium. The

92OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

same workers were not necessarily common to each sampling period. The highest serum PFOA concentrations have been reported in workers at3M’s Cottage Grove facility in Minnesota. The mean PFOA concentration for Cottage Grove workers in 2002 was 4.3 ug/ml (range, 0.07 – 32.6) (Olsen, et al., 2003a). A summary of the occupational da ta collected at the 3M facili ties are presented in Table 21.

Table 21 Summary of 3M Biomonitoring Data for Occupational Exposures (serum concentrations, μg/ml)*

Plant Arithmetic Mean Range Geometric Mean 95% CI of Geometric Mean Reference
Cottage Grove 2002 (n=38) 2000 (n = 148) Male (n = 131) Female (n = 17) 1997 (n =74) 1995 (n = 80) 1993 (n = 111) 4.3
NR
4.51
0.85
6.4
6.8
5.0 0.07 - 32.6
NR
.007-92.03
.04 - 4.73
0.1 - 81.3
0 - 114.1
0 - 80.0 1.7
NR
0.85
0.42
NR
NR
NR 1.02-2.72
NR
0.64- 1.22
0.23 - 0.79
NR
NR
NR Olsen, et al., 2003a Olsen, et al., 2003f

Olsen, et al., 2000 Olsen, et al., 2000 Olsen, et al., 2000
Decatur 2002 (n =54) 2000 (n = 263) Male (n = 215) Female ( n = 48) 1998 (n = 126) Male (n =102) Female (n=24) 1997 (n = 84) (males only) 1995 (n = 90) (males only) 1.497
1.78
1.90
1.23
1.54
1.735
0.691
1.40

1.72 0.025-4.81
0.04-12.70
0.04-12.70
NR
0.02 - 6.76
0.02 - 6.76
NR
NR

NR 0.713
1.13
NR
NR
0.90
1.142
0.326
NR

NR 0.483-1.055
0.99 - 1.30
NR
NR
0.72 -1.12
NR
NR
NR

NR Olsen, et al., 2003b Olsen, et al., 2001a Olsen et al., 2003e Olsen et al., 2003e Olsen, et al., 1999

Olsen, et al., 1998c

Olsen, et al., 1998c

Antwerp 2003 (n = 30) 2000 (n = 258) 1995 (n = 93) 2.63
0.84
1.13 0.92 – 5.69
0.01 – 7.04
0.00 – 13.2 2.35
0.33
* 1.96 – 2.82
0.27 – 0.40
* Olsen et al., 2003c Olsen, et al., 2001c

* μg/ml = ppm NR not reported

2.3.2.2 DuPont Occupational Data Dupont has been measuring PFOA in worker s’ blood serum since 1981 (Dupont, 2001a, 2001b). Prior to 1981, total blood fluoride levels were reported. All of Dupont’s operations in the U.S. that use PFOA with significant exposure potential are conc entrated at the Washi ngton Works plant in Washington, West Virginia. PFOA is used as a fl uoropolymer reaction aid at this plant. The data

93OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

presented in Table 22 reflect serum concentrati ons of volunteer workers in the plant who had potential PFOA exposure. With the exception of the 2000 sampling period, the same workers were common to all of the other sampling periods, withadditional workers participating in 1995. The data available at this time are limited to what is presented here.

Table 22 Summary of DuPont Occ upational PFOA Biomonitoring (serum concentrations, μg/ml)*

Year Arithmetic Mean Range Reference

Washington Works 2004 (n = 259) 2000 (n= 72) 1995 (n = 80) 1989-90 (n =22 ) 1985 (n = 22) 1984 (n = 19) 0.494 a
1.53
1.56
1.96
2.34
3.21 0.0174 - 9.55
0.02 - 9.0
0.12 - 4.5
0.06 - 11
0.06 - 18
0.07 - 24 DuPont, 2004 DuPont, 2001a DuPpont, 2001b DuPont, 2001b DuPont, 2001b DuPont, 2001b * μg/ml = ppm a median (Teflon workers only)

2.3.3 General Population Exposures Limited data on PFOA concentrations in the ge neral population of various countries throughout the world are currently available. Most of the data pr esented here are the first to be collected on human populations, and most of the sample sizes are fairly small. Data on both pooled a nd individual samples are available. PFOA has been detected in human blood (plasma, serum and whole blood), liver, seminal plasma, and breast milk. Based on the data currently available, the ranges of PFOA concentrations indicate that the interind ividual variability can be quite large in some populations and that human serum concentrations can vary between countries, as well. In addition, preliminary US reports indicate that individuals living near a f acility that uses PFOA ma y have much higher PFOA serum concentrations than the levels previously reported for US populations (Emmett et al., 2006). Very little data on trends in human populations are currently available. Blood samples co llected in western Maryland in 1974 and 1989 (n = 178 for each time period) indicate that median PFOA concentrations have more than doubled over this time period (Olsen, et al., 2005). However, other samples from different study participants (n=108) co llected in the same region in 2001 did not indicate that the geometric mean was statistically significantly diffe rent between 1989 and 2001. On the other hand, a recent report has show n that PFOA serum concentrations in Japanese residents have steadily increased over the past 25 years (Harada, et al., 2004). Pooled Serum Data in the General Population Pooled serum PFOA data from US residents have been reported for 1998, 1999, 1990-1998, 2002 and 2001-2002. Reported means and ranges for pooled blood samples collected in the US in 1998 (n = 18 lots, 340-680 donors) and 1999 (n = 35 lots ) were 3 and 17 ppb (μg /l), respectively, with ranges of 1-13 ppb and 12-22 ppb (3M 1999a, 1999b). In another study, 23 archived serum pools consisting of samples from various US locations were analyzed for PFOA and other perfluorinated co mpounds. The samples were collected from 1990- 1998 and in 2002 (Calafat, et al., 2005). Fifteen pools were prep ared from the 1990-1998 pools and were comprised of serum from 40 – 200 resident s of the southeastern US. One of the 2002 pools

94OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

95 was comprised of serum from 15 Louisville, KY resi dents; another 2002 pool was from 15 residents of Philadelphia, Memphis, and Miami. Six seru m pools of samples collected from 1999-2001 were each made up of serum of 6-8 Seattle residents. In 2003, blood samples were also collected from 44 residents of Trujillo, Peru (25 me n, 19 pregnant women) and analyzed for perfluorinated chemicals. PFOA was detected in 100% of the US samples, wh ile it was only detected in 25% of the samples collected in Peru. In th e US pooled samples, the geometric mean of PFOA was 9.6 μg/l (range 2.8 – 23.7) (median 11.6). When the samples were anal yzed by geographic region, the mean PFOA concentration was 5.8 μg/l (s.d. 0.6) for the Seattle samples and 14.1 μg/l (s.d. 7.6) for the pooled samples collected in the eastern U.S. The 95th per centile serum concentration was 0.3 μg/l for the samples collected in Peru. A recent study comprised of 54 pooled serum samp les collected from 1833 participants 12 years and older of the 2001-2002 NHANES showed mean PFOA serum concentrations highest for nonHispanic whites, followed by non-Hispanic blacks, and Mexican Americans. Females had lower mean PFOA concentrations than males, and adol escents aged 12-19 had highe r concentrations than adults 20 years of age and older (Calafat AM, 2006). Individual Serum Data in the General Population Results of analyses on individual human biological samples have beenreported in the literature for several countries. With the excepti on of 3 studies 3M conducted on fa irly large cohorts in the U.S., but that are now 5-10 years old, mo st of the sample sizes are sma ll. However, based on these data which are available thus far, it is indicated that there are geographic diffe rences both between and within countries in exposure to PFOA. Data currently available on worldwide PFOA concentr ations in various biologi cal matrices in individual samples in human populations are presente d in Table 23. The concentrations reported for whole blood have been converted to serum c oncentrations by doubling the whole blood levels (Kannan, et al., 2004). Based on a recent analysis of 12 production workers occupationally exposed to PFOA, plasma and serum concentrations were ve ry similar in each individual but were approximately 51% higher than whole blood concentrations; therefore, using a factor of 2 to estimate serum concentrations from whole blood seems reasonable (E hresman et al., 2005a). In addition, analytical laboratory techniques are not reflected in Table 23, and various methods can be used to analyze perfluorochemicals. A recent interlaboratory study on perfluorinated compounds has concluded that the overall concordance between laboratories for PFOA analysis is “satisfact ory”; however, the variability for PFOA analysis approximated as much as 50% (Van Leeuwen, et al., 2005). Details of the three largest biomonitoring studies on perfluorin ated chemicals are provided following Table 23. Table 23 Summary of PFOA Individual Biom onitoring Data in Various Countries (serum concentrations, ng/ml)*

Location Age Matrix Year of Collection N Average 1,2,3 Range Reference

American Red Cross Blood Banks(6 US locations) adults (20- 69) serum 2000 645 5.6 1 1.9 – 52.3 Olsen et al., 2003

Seattle, WA elderly (65- 96) serum 2000 238 4.2 2 1.4 – 16.7 Olsen et al., 2004

23 US states children (2- 12) serum 1995 598 5.6 1 1.9 – 56.1 Olsen et al., 2004OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

96 Location Age Matrix Year of Collection N Average 1,2,3 Range Reference

Central Michigan female adults
male adults serum 2000 46
29 4.7 1

5.7 1 <3 - 7.3
<3 – 14.7 Kannan et al., 2004

Murray, KY female adults
male adults whole blood 4 2002 11
19 23 1

41.6 1 15 - 39
11 - 88 Kannan et al., 2004

New York, NY unknown plasma 2002 70 27.5 1 14 - 56 Kannan et al., 2004

Atlanta, GA adults serum 2003 20 4.89 1 0.2 – 10.4 Kuklenyik et al., 2004 Samples frombiological supply company (US) unknown serum unknown 65 6.4 <5 - 35.2 Hansen et al., 2001

Sera and liver tissue from IAM 5 – 74 years serum liver NR 23 pairs -- 90% < LLOQ <0.0030 - 0.0070
<0. – 0.0147 Olsen et al., 2003

Hagerstown, MD adults serum plasma 1974
1989 178
178 2.1 2

5.5 2 1.9 - 2.25
5.2 - 5.95 Olsen et al., 2005
Ottawa, Ontario,Gatineau, Quebec adults serum 2002 56 3.4 3 <1.2 – 7.2 Kubwabo et al., 2004

Cartagena, Columbia female adults
male adults whole blood 2003 25
31 6.1
6.2 3.7 – 9.2
3.9 – 12.2 Kannan et al., 2004

Gdansk, Poland female adults
male adults whole blood 2003 15
10 21.9
20.5 9.7 - 34
11 – 40 Kannan et al., 2004

Flanders, Wallonia, Belgium female adults
male adults plasma 1998, 2000 4
16 4.1
5.0 < 1 – 7.6
1.1 – 13 Kannan et al., 2004

Coimbatore, India female adults
male adults serum 2000 11
34 < 3 (0% detected) 3.5 (3%) < 3
<3 – 3.5 Kannan et al., 2004

Daegu, Korea female adults
male adults whole blood 2003 25
25 88.1 (19%) 35.5 (25%) 15 – 256
< 15 – 71.4 Kannan et al., 2004

Colombo, Talawakele,
Haldummulla Sri Lanka male adults serum seminal plasma 2003 30
0.323 6.38
<0.072 – 0.621 0.23 – 23.5 Guruge, et al., 2005

Kyoto, Japan male adults
female adults serum 2004 20
20 12.96-12.41
7.89-12.63 5.2 – 15.8
7.6 - 20.6 Harada, et al., 2005OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

97 Location Age Matrix Year of Collection N Average 1,2,3 Range Reference

Kyoto, Japan female adults
male adults serum 2003 20
14 7.1 2

12.4 2 2.8 – 12.4 Harada, et al., 2004

Miyagi, Japan female adults
male adults
female adults serum

serum

serum 2003
2003
1977 23
32
39 2.8 2

3.3 2

0.2 2 NR
NR
NR Harada, et al., 2004

Yokohama, Tsukuba Japan female adults
male adults serum 2002 13
25 12.3 (8%) <6.8 (0%) < 6.8 – 12.3
< 6.8 Kannan et al., 2004

Akita, Japan female adults
male adults
female adults
female adults
male adults serum

serum

serum

serum

serum 2003
2003
1995
1991
1991 50
66
40
60
16 2.5 2

3.4 2

1.9 2

1.8 2

2.2 2 NR
NR
NR
NR
NR Harada, et al., 2004

Kyoto, Japan

10 locations Adults

Female adults
male adults serum 1983
1999

2003 100

200 2.2 2

9.7 2

3.5 2

4.2 2 0.4 – 52.2 Saito, et al., 2005

Sweden female adults serum NR 108 2.0 1 NR (SD = 0.8) Holmstrom et al., 2005 Zhoushan, China female adults breast milk 2004 19 NR 47 – 210 5 So et al., 2006bGermany, environmental specimen bank male, female
20-29 plasma 2004 16 6.5 4.1-33.2 UBA, 2004

* ng/ml = ppb; NR – not reported 1 Mean 2 Geometric Mean 3 16 samples above LOQ 4 whole blood results are estimates of serum concentrations 5 ng/L (ppt)

Individual blood samples from 3 different age populations were analyzed for PFOA and other fluorochemicals using high-pressure liquid chromatography/electrospr ay tandem mass spectrometry (HPLC/ESMS) (Olsen et al., 2002a, 2002b, 2002c). Th e 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. Overal l, the PFOA geometric means were similar across all 3 populations (4.6, 4.2 , and 4.9 ng/ml, respectively) . The geometric means and 95% tolerance lim-OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

its (the exposure below which 95% of the populat ion is 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 othe r factor. It should be noted that PFOS and PFOA were highly correlated in all thr ee studies (r = .63, r = .70, and r = .75) and that PFOA did not meet the criteria for a log normal distri bution based on the Shapiro-Wilk te st in any of the studies. The authors suggest that it may be due to the greate r proportion of subjects withvalues less than the lower limit of quantitation (LLOQ); however, only 12 of the 1481 total samples were below the LLOQ. In those instances where a sample was measured below the LLOQ, the midpoint between zero and the LLOQ was used for ca lculation of the geometric mea n. 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 f actors known for each donor were age, gender, and location. The geometric mean serum PFOA level was 4.6 ng/ml. The range was <lower limit of quantitation (1.9 ng/ml) to 52.3 ng/ml. Only 2 samples were less than the LLOQ. Males ha d significantly higher (p < .05) geometric mean PFOA levels than female s (4.9 ng/ml vs. 4.2 ng/ml). Age was not an important predictor of adult serum fluorochemical concentrations. When stra tified by geographic location, the highest geometric mean for PFOA was in the samples from Ch arlotte, NC (6.3 ng/ml, range: 2.1 – 2.9 ng/ml) and the lowest from Portland (3.6 ng/ml, range: 2.1 – 16.7 ng/ml). The highest individual value was reporte d in Hagerstown (52.3 ng/ml). Serum PFOA levels were reported for 238 (118 male s and 120 females) elderly volunteers in Seatle participating in a study designed to examine cognitive function in adul ts aged 65-96 (Olsen et al., 2002b, 2004a). Age, gender and number of years’ resi dence 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 ng/ml. The range was 1.4 – 16.7 ng/ml. Only 5 samples were less than the LLO Q of 1.4 ng/ml. There was no signi ficant (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. 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 (O lsen et al., 2002c, 2004b) . The samples were collected in 1994-1995 from children residing in 23 stat es and the District of Columbia. PFOA did not meet the criteria for a log normal distribution base d 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 5 samples were less than the LLOQ of 1.9 ng/ ml. The geometric mean of PFOA for all of the participants was 0.0049 ng/ml, and the range was 1.9 to 56.1 ng/ml. Male children had significantly (p<.01) higher geometric mean serum PFOA levels than females: 5.2 ng/ml and 4.7 ng/ml, respectively. In simple linear re gression analyses, age was significan tly (p < .05) negatively associated with PFOA in both males and females. When strati fied by age, the geometric mean of PFOA was highest at age 4 (5.7 ng/ml) and lo west at age 12 (3.5 ng/ml). Alt hough the data were not reported, a graphical presentation of log PFOA levels for eachstate by gender looked si milar across the states. However, it is difficult to inte rpret these data without analyzing them and the sample sizes were limited for each gender/location subgroup. PFOS and PFO A were highly correlated (r = .70) in this

98OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

study. PFOA and PFHS (perfluorohexanesulfonate) were also correlated, although not as strongly (r = .48).

The above 3 studies indicate sim ilar geometric means and ranges of PFOA among sampled adults, children, and an elderly population. However, anunexpected finding was the level of PFHS and M570 (N-methyl perfluorooctanesulfonamidoacetate) in children (Olsen, 2002c, 2004b). 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 differentexposure pattern in children. In another study, the PFOA concentration was anal yzed in human sera and liver samples using HPLC/ESMS (Olsen et al., 2001d). Thirty-one donor samples were obtained from 16 males and 15 females. The average age of the male donors wa s 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 donor s contributed liv er tissue only and 1 donor contributed serum only. Resulting se rum 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 calculated as 0.0031 ng/ml with a geometric mean of 0.0025 ng/ml. No liver to serum ratios we re provided because more than 90 % of the individual liver samples were <LOQ.

Conclusions
Data are available on PFOA blood serum levels in both occupationally- and non-occupationallyexposed populations. Although the samp le sizes have been relatively sm all, and it is currently difficult to compare or interpret PFOA concentrations without nationally repr esentative samples and standard analytical methods, PFO A has been detected in sample s of human blood (plasma, serum and whole blood), liver, seminal plasma, and breast milk from several countries throughout the world, including the US, Canada, Columbia, Poland, Belgium, Indi a, Korea, Sri Lanka, Japan, Sweden, and Germany. Mean blood serum concentra tions in most of these countries seem to be generally similar to those in the US, where mostof the human biomonitoring data have been collected thus far. However, preliminary US reports i ndicate that individuals livi ng near a facility that uses PFOA have much higher PFOA serum concentrations than the levels previously reported for US populations.

2.3.4 Consumer Exposure In a collaboration between The Norwegian, Swedish and Danish Soci eties for Nature Conservation, textile samples were taken from a range of all-w eather jackets for children in the Nordic Market. Alltogether 6 jackets from 5 diff erent brands were collected from shops or warehouses in Sweden and Norway. The jackets investigated came from Norwegian, Swedish and foreign manufacturers. Common to the textile was that they were so-calle d breathing synthetic textiles, with the exception of one cotton anorak material. Samples (10 x 10 cm; 1.04-1.91 g) were extracted with ethyl acetateand methanol and analysed via GC/MS and HPLC/MS. PFOA was detected in concentrations ranging from 0.80 to 24.6 μg/m2 and 6.1 to 107 ng/g. Acco rding to the analysis re sults the cotton anorak contained 127 μg/m2 and 416 ng/g PFOA (SNF 2006). According to Kuklenyik et al. (2004) PFOA is us ed primarily in the production of fluoropolymers, such as poly(tetrafluoroethylene) (PTFE) and poly(vinylidine fluoride), and fluoroelastomers. These polymers have numerous uses in many industrial and consumer products, including soil-, stain-, grease-, and water-resistant coatings on textiles and carpet, personal care products, and nonstick

99OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

coatings on cookware, and uses in the automotiv e, mechanical, aerospace, chemical, electrical, medical, and building and construction industries. Consumer exposure to PFOA due to impurities in the finished product can not be excluded. Washburn et al. (2005) investigated the potent ial exposure towards PFO(A) during consumer use of certain articles. Certain consum er articles are manufactured with either fluorotelomer s or fluoropolymers. Fluorotelome r-based products are typically manufactured through telomerization of tetrafluor oethylene with perfluor oethyl iodide. Trace amounts of PFO may be generated during this process as an unint ended by-product. Fluorotelomerbased products are used to treat a variety of cons umer articles (e.g., apparel and carpeting), primarily to impart stain and soil resistance. Fluor opolymers are films (e.g., on nonstick cookware) or membranes (e.g., in outerwear). Th e ammonium perfluorooctanoate (A PFO) is an essential processing aid in the formulation of such fluoropolymer s; residual PFO(A) may be present in fluoropolymer films and membranes used in manufacturing certain consumer articles. The authors analysed (LC/MS/MS) fluorotelomer-based product formulati ons and finished consumer articles manufactured with fluoropolymers or fluorotelomer-basedproducts for the occurrence of PFOA. The analytical test results are summarized in Table 24. As th e analytical test results are based on the extraction of new articles the values obtained are representing worst case situations. Although PFO(A) was not detected in the fluorotel omer-based product applied to th e upholstery PFO(A) was detected in extraction tests on treated upho lstery. According to the opinion of the authors this may result from unrelated background sources of PFO(A) on upholstery. A simple one-compartment model was developed to estimate contributions of potentia l consumer exposures to PFO concentrations in serum. Based on a simple compartmental model es timation, the hypothetical ag gregate exposures to the consumer articles investigated correspond to serum concentrations ranging from approximately 0.05 to 0.25 ppb for the residential adolescent and adult and from about 0.05 to 0.25 ppb for the professional and therefore less th an the current quantification limit for PFO(A) in serum (ca.0.5 ppb). In adults of th e general U.S. population the geometric mean of PFO(A) is repo rted to be ca. 5 ppb. Based on this estimation an explanation for the PFO(A) concentrations found in humans can not be given.

Table 24 PFOA concentration in fluoro telomer and fluoropolymer articles

Articles Concentration in Fluorotelomer product (mg PFO(A)/l) Calculated total concentration in finished article (mg PFO(A)/kg) Extraction of finished article (ng PFO(A)/cm2 article)* Fluorotelomer-Containing Articles mill-treated carpeting (n>60) 30-80 0.2-0.6 <0.2 (l.o.d.) to 23

carpet-care solution-treated carpeting (n=14) 1-50 0.2-2 28-50

treated apparel (n >100) <1 (l.o.d.) to 40 <0.02 (l.o.d.) to 1.4 <0.01 (l.o.d.) to 12 treated upholstery (n=3) <1 (l.o.d.) <0.034 (l.o.d.) 0.4-4
treated home textiles <1(l.o.d.)) to 40 <0.02(l.o.d.) to 1.4 Not tested
treated technical textiles <1(l.o .d.) <0.034(l.o.d.) Not tested
treated nonwoven medical garments (n=6) <1(l.o.d.) <0.034(l.o.d.) <0.02(l.o.d.)

stone, tile, and wood sealants <1(l.o.d.) <0.1(l.o.d.) Not tested

industrial floor waxes and

wax removers 5-120 0.0005-0.06 Not tested

100OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

101 Articles Concentration in Fluorotelomer product (mg PFO(A)/l) Calculated total concentration in finished article (mg PFO(A)/kg) Extraction of finished article (ng PFO(A)/cm2 article)* latex paint 50-150 0.02-0.08 Not tested
home and office cleaners 50-150 0.005-0.05 Not tested Fluoropolymer-Cont aining Articles membranes for apparel (n=20) Not applicable Not applicable 0.008-0.07

non-stick cookware (n=40)** Not applicable Not applicable <0.1(l.o.d.)

thread seal tape (n=20) Not applicable Not applicable 0.02-0.08 * simulation of human exposure (extractants: water, saliva, perspiration) ** manufacture of nonstick cookware includes a high temperature step which reduces residual PFO(A) prior to article use by consumers

According to a study performed in contract for a journal for consum er protection (without peer review process), PFOA was detected in all 5 jacket s tested at concentrations ranging from 0.08 to 0.60 mg/kg. In 2 out of 3 impregnation sprays 0.2 4 and 0.48 mg/kg PFOA were detected (Öko-Test 2005). Samples were taken from the outside of jackets, extracted wi th methanol and analysed using HPLC-MS/MS negative uib elec trospray (Eurofins, 2005). Larsen et al. (2005) determined perfluorooctanoic acid in polytetrafluoroethylene polymers. In one polymer which was an intermediate of a comme rcial lot a PFOA concentration of 1040 ppb was detected, in the other polymer obtained from a commercial lot the corresponding PFOA concentration was 59 ppb. Based on the resu lts the authors concluded that the fi nishing step significantly reduced PFOA concentration in the commercial polymer. Powley et al. (2005a) developed an analytical met hod for the determination of PFOA released from the surface of commercial cookwar e. The cookware was extracted with water and water/ethanol mixtures at 100 and 125°C and the extracts analy zed via LC/MS/MS. PFOA could not be detected in any of the samples at a detection limit of 100 pg/cm -2

. Powley et al. (2005b) examined fry pans purchased from retail for extractable PFOA. The pans were heated with water for 30 minutes. The wate r was cooled and analysed via HPLC/MS/MS. Further more fry pans were cut into pieces and plac ed in a pressurized fluid extraction apparatus and extracted with etanol/water (1:9 and 19:1) to s imulate watery and acidic foods and fatty or oily foods. None of the fluoropolymer-treated pans showed any detectable levels of PFOA. Ellis et al. (2001) investigated the thermal d ecomposition of fluoropolymer s at 360 and 500°C. The thermolysis of PTFE produced perfluoroalkyl aci ds with C atoms ranging from 3-14. The amount of PFOA resulting from the thermolysis of the fluoropol ymer is not stated; the sum of PFCs (C3-C14) is given to >0.01%. PFOA was not detected as th ermal degradation product of other fluoropolymers (CPTFE: chloropolytrifluoroethylene, ECTFE: ethylene-chlorotrifluoro-ethylen, PFEPE). Boulanger et al. (2005) analyseda 1994 Scotchgard sample (0.832 g was sprayed into a HDPE bottle, dissolved in methanol and analyzed usi ng HPLC/MS/MS). The PFOA concentration found corresponded to 1.34 x 10 -3 % (=82% of perfluorinated compounds). The authors assume that the presence of perfluorinated compounds (e.g. PFOA) was the result of residual monomers in the formula. Based on the analysis of human blood sera and se minal plasma collected from volunteers in Sri Lanka Guruge et al. (2005) concluded that the gr eater exposure levels found in Colombo are indica-OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

tive of sources from various industries and frequent use of products containing PFCs such as paper, packing products, carpet spray, st ain-resistent textiles, cosmetics, electronics and fire fighting foams. In the frame of a Multi-City Study, a market basket sampling was conducted. In a total over 200 samples taken from green beans, apples, pork mu scle, cow milk, chicken muscle, chicken eggs, bread, hot dogs, catfish, and ground beef were analysed for PFOA. Measurable quantities of PFOA, ranging up to 2.35 ng/g, were found in two ground b eef samples, two bread samples, two apple samples, and one green bean sample (3M 2001c) Begley et al. (2005) analysed PTFE cookware for PFOA content. Food cooked using PTFE cookware was also analysed for PFOA. Results of mi gration tests show mg/kg amounts of perfluoro paper additives/coatings transfer to food oil. An alysis of PTFE cookware shows residual amounts of PFOA in the low μ g/kg range. PFOA was detected in microwave popcorn bag paper in concentrations up to 300 μ g/kg. Kubwabo et al. (2005) investig ated house dust samples for the oc currence of perfluoro chemicals (PFOS, PFOA, perfluorobutane su lfonate (PFBS), perfluorohexane sulfonate (PFHS) and perfluorooctane sulfonamide (PFOSA). The data revealed a correlation between the concentrations of PFCs and the percentage of carpeting in the house; older houses tended to have less carpeting, hence lower levels of these perfl uorinated compounds in their dust.

3 HUMAN HEALTH HAZARDS

3.1 Metabolism and Pharmacokinetics

3.1.1 Metabolism and Pharmacoki netic Studies in Humans

Little is known about the metabolism and pharmacokinetics of PFOA in humans. One report notes the presence of PFOA in the cord blood of some pregnant workers suggesting that PFOA can cross the placenta (U.S. EPA, 2003). However, PFOA was notdetected in cord blood in a study of 15 pairs of maternal and cord blood samples in non-occupationally e xposed women and infants in Japan (Inoue et al., 2004). PFOA has also been detected in seminal plasma in tea workers in Sri Lanka (Guruge et al., 2005) and in very lo w concentrations in breast milk in a small sample of women in China (So et al., 2005).
There are limited data on the half-life of PFOA in humans. With the ex ception of a 1980 study in which total organic fluorine in blood serum wa s measured in one worker (Ubel et al., 1980), no other data were available until June 2000. 3M has conducted a half-life study on 26 retired fluorochemical production workers from their Decatur, Alabama (n = 24) and Cottage Grove, Minnesota (n = 3) plants in which serum samples were dr awn every 6 months over a 5-year period. Two interim reports describing the results thus far were submitted to EPA (Burris et al., 2000; Burris et al., 2002). Final results have not yet been published, how ever, preliminary data have been presented which indicate that the mean se rum elimination half-life of PF OA in these workers was 3.8 years (1378 days, 95% CI, 1131-1624 days) and the median was 3.5 years (O lsen et al., 2005). The range was 1.5 - 9.1 years (561-3334 days). Some of the details of this study are provided below. Blood was collected from the subjects between 1998 and 2004, depending on the facility at which the subject had worked. Responses on questionnaire s determined whether any of the retirees had any occupational exposures after re tirement. One subject was exclude d from the analysis based on his response. The average number of years that this sample worked was 31 (range 20-36 years) and

102OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

they had been retired an averag e of 2.6 years at study initiation (range 0.4 – 11.5 years). The mean age of these retirees was 61 years (range 55-75) at the beginning of the study.
The initial mean serum PFOA conc entration of all of the subj ects was 0.691 μg/ml (range, 0.072 – 5.1 μg/ml). At the completion of the study, the m ean PFOA concentration was 0.262 μg/ml (range, 0.017 – 2.435 μg/ml). Two of the reti rees died during the study period ; therefore, th ey were only followed for 4.2 years. No association was report ed between the serum elimination half-life and with initial PFOA concentrations, age or sex of retir ee, the number of years retired or worked at the production facility, or medication use or health conditions. Harada et al. (2005) studied th e relationship between age, sex and serum PFOA concentration in residents of Kyoto, Japan. They found that female s in the 20-50-year-old age group (a ll with regular menstrual cycles) had serum PFOA concentrations th at were significantly lower than those in females over age 50 (all postmenopausal). Mean serum PFOA concen tration in the younger females was 7.89 ± 3.61 ng/ml versus 12.63 ± 2.42 ng/ml in the older females. This age difference in serum PFOA concentrations was not seen in males, a nd serum PFOA concentrations in males were comparable to those of the older females. Harada et al . (2005) also estimated the renal clearance rate of PFOA in humans and found it to be only about 0.001% of the glomer ular filtration rate, indicating the absence of active excretion in human kidne ys. There was no significant difference in renal clearance of PFOA with respect to sex orage group, and the mean value was 0.03 ± 0.013 ml/day/kg.

3.1.2 Metabolism and Pharmacokinetic Studies in Non-Human Primates Butenhoff et al. (2004b) studied th e fate of PFOA in cynomolgus m onkeys in a six-month oral exposure study. Groups of four to six male monkeys each were administered APFO daily via oral capsule at dose rates of 0, 3, 10, or 20 mg/kg (30 mg /kg for 12 days, reduced to 20 mg/kg when dosing reinitiated on test day 22) for six months. Serum, urine and fecal samples were collected at twoweek intervals and were analyzed for PFOA concentr ations. Liver samples were collected at time of sacrifice. Serum concentrations reached steady-state leve ls within four week s in all dose groups. Steady-state concentrations of PFOA in serum were 81 ± 40, 99 ± 50, and 156 ± 103 μg/ml for the 3, 10, and 20 mg/kg dose groups, respectively. The m ean serum concentration of PFOA in control monkeys was 0.134 ± 0.113 μg/ml. Elimination of PF OA from serum after cessation of dosing was monitored in recovery monkeys from the 10 and 20 mg/kg dose groups. For the two 10 mg/kg recovery monkeys, serum PFOA elimination half-life was 19.5 (R2 =0.98) days and indicated firstorder elimination kinetics. For three 20 mg/kg monkeys, serum PFOA elimination half-life was 20.8 days (R2 =0.82) and also indicated fi rst-order elimination kinetics . Urine PFOA concentrations reached steady-state after 4 weeks at 53 ± 25, 166 ± 83, and 181 ± 100 μg/ml in the 3, 10, and 20 mg/kg dose groups, respectively. Within two weeks of recovery, urine PFOA concentrations were <1% of the previous value measured during treatm ent and decreased slowly thereafter. Fecal PFOA concentrations were 6.8 ± 5.3, 28 ± 20, and 50 ± 33 μg/ml in the 3, 10, and 20 mg/kg dose groups, respectively. Within two weeks of recovery, fecal PFOA concentr ations dropped to less than 10% of the last value during treatment, and then declin ed slowly. Liver PFOA con centrations at terminal sacrifice in the 3 mg/kg and 10 mg/kg dose groups were similar and ranged from 6.29 to 21.9 μg/g. Liver PFOA concentrations in the two 20 mg/k g monkeys were 16.0 and 83.3 μg/g. Liver PFOA concentrations in two 10 mg/kg-day monkeys at the end of the recoveryperiod were 0.08 and 0.15 μg/g. Butenhoff et al. (2004b) also conducted an intrav enous pharmacokinetic study in cynomolgus monkeys. Three male and three female monkeys were administered a single iv dose at 10 mg/kg of potassium PFOA. Blood samples were collected prior to dose administration; at 0.5, 2, 4, 8, and 24 hr; and on Days 2, 4, 7, 11, 14, 21, 28, 57, 79, 87, and 123 post-dose. Urine was collected in metabo103OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

lism cages for 24-hr intervals on the following days : prior to dose administration, Day 1 (0-24 hr post-dose), and on Days 2, 7, 14, 21, and 28. Serum PFOA concentr ations at 0.5 hr post-dose ranged from 91 to 97 μg/ml in the male monkeys a nd from 89 to 96 μg/ml in the females. At 48 hr, serum PFOA concentrations were between 49 and 66 μg/ml in the males and 40 and 62 μg/ml in the females. After 48 hours, serum PFOA concentrations in two of the three male s decreased at a faster rate than that observed in the third male monkey and the three female monkeys. On day 28, serum PFOA concentrations ranged from 1.9 to 27 μg/ml inthe male monkeys and from 7.1 to 34 μg/ml in the females. Substantial amounts of PFOA were excret ed in the urine, ranging from 3 to 19 % of the dose on Day 1 and from 1.3 to 14.1 % on Day 7. Urinary excretion of PFOA was prolonged, evidenced by 0.1 to 1 % of the dose in urine on Da y 28. Pharmacokinetic parameters were calculated from serum PFOA concentrations using a non-compartmental analys is (Table 25). AUC0-infinity values ranged from 571 to 2501 μg/•day/ml (mean : 1235 μg/•day/ml) for male monkeys and from 1094 to 3224 μg/•day/ml (mean: 2293 μg/•day/ml) for female monkeys. The terminal half-lives of PFOA in serum were 13.6, 13.7, and 35.3 days in the three male monkeys and 26.8, 29.3, and 41.7 days in the three females. The total body clear ances were 4.0, 15.8, and 17.5 ml/day/kg for males and 3.1, 3.9, and 9.1 ml/day/kg for females. Volumes of distribution of PFOA at steady state were 168, 184, and 192 ml/kg in males and 133, 190, and 270 m l/kg in females, suggesting distribution primarily in the ex tracellular space.

Table 25 Pharmacokinetic Parameters Calc ulated from Serum Concentrations of PFOA in Cynomolgus Monkeys

Males Females
Parameter 2052 2054 2211 Mean S.D. 2058 2059 2061 Mean S.D.
Cmax (μg/ml) a 101 98.4 91.6 97.0 4.9 105 93.0 103 100 6.0

T 1⁄2 (day) b 35.3 13.7 13.6 20.9 12.5 26.8 41.7 29.3 32.6 8.0
AUC 0-last (μg•day/ml) c 1999 569 633 1067 808 3083 2314 1056 2151 1023

AUC 0-infinity (μg•day/ml) d 2501 571 634 1235 1097 3224 2560 1094 2293 1090

Cl (ml/day/kg) e 4.0 17.5 15.8 12.4 7.7 3.1 3.9 9.1 5.4 3.3
Vd ss (ml/kg) f 192 168 184 181 12 133 190 270 198 69 (from Butenhoff et al., 2004b) a Maximum concentration in serum b Half-life of the terminal elimination phase c Area under the serum concentration versus time curve calculated from 0 to the last time point d Area under the serum concentration versus time curve calculated from 0 to infinity e Total body clearance f Volume of distribution at steady state

104OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

3.1.3 Metabolism and Pharmacokine tic Studies in Adult Rats

3.1.3.1 Absorption Studies

Studies in rats have shown that PFOA is absorb ed following oral, inhalati on, and dermal exposures.

3.1.3.1.1 Oral Exposure Gibson and Johnson (1979) administered a single dose of 14 C-APFO averaging 11.4 mg/kg by gavage to groups of 3 male 10-week old CD rats. Twenty-four hours after admi nistration, at least 93% of the total carbon-14 was absorbed; the eliminati on half-life of carbon-14 from the plasma was 4.8 days. Ophaug and Singer (1980) administered 2 ml of an aqueous solution of 2 mg PFOA to female Holtzman rats. Ionic fluoride, nonionic fluorine, and total fluorine were measured. Within 4.5 hr, 37% of the total fluorine in the administered dose was recovered in the urine. The quantity of nonionic fluorine recovered in the urine increased to 61%, 76% and 89% at 8, 24 and 96 hr, respectively, after administration. Within4.5 hr, serum from treated rats had a nonionic fluorine level of 13.6 ppm. 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. Despite the large decreas e in nonionic fluorine in the serum, the ionic fluoride level was only 0.03 ppm an d remained at that level throughoutthe experiment. Prior to administration of PFOA, the ionic and nonionic fl uorine background levels in serum were 0.032 and 0.07 ppm, respectively. The authors concluded that PFOA is rapidly absorbed from the gastrointestinal tract and then rapidly cleared from the serum.

3.1.3.1.2 Inhalation Exposure Hinderliter (2003) measured the serum concentrati ons of PFOA following singl e and repeated inhalation exposures in Sprague-Dawley rats. For the single exposure study, ma le and female rats (3/sex/group) were exposed to a single nose-onlyexposure of an aerosol of 0, 1, 10, or 25 mg/m 3 PFOA. Preliminary range-finding studies demonstrated that aerosol sizes were 1.8 - 2.0 μm mass median aerodynamic diameter (MMAD) with geometric standard deviations ranging from 1.9 - 2.1 μm. Blood samples were collected pre-exposure, at 0.5, 1, 3 and 6 hours during exposure, and at 1, 3, 6, 12, 18 and 24 hours after exposure. Plasma wa s analyzed by LC-MS. PFOA plasma concentrations were proportional to the inha lation exposure concentrations. The male Cmax values were approximately 2-3 times higher than the female Cmax . The female Cmax occurred approximately one hour after the exposure period, whil e the male Cmax occurred from the end of the exposure period up to six hours after exposure. In females, the el imination of PFOA was rapid at all exposure levels, and by 12 hours after exposure the plasma levels ha d dropped below the analyt ical limit of quantitation (0.1 μg/ml). In males, the plasma eliminatio n was much slower, and at24 hours after exposure, the plasma concentrations were approximately 90% of the peak concentrati ons at all exposure levels. In the repeated dose study, Hinderliter (2003) expos ed male and female rats (5/sex/group) to the same aerosol concentrations of PFOA for 6 hrs/da y, 5 days/week for 3 week s. Blood was collected immediately before and after the daily exposure pe riod three days per week. The aerosol sizes were 1.3 - 1.9 μm MMAD with geometric standard deviat ions of 1.5 - 2.1. PFOA plasma concentrations were proportional to the inhalation exposure concentrations, and repe ated exposures produced little plasma carryover in females, but significant carryove r in males. Male rats apparently reached steady state plasma levels by three week s with plasma concentrations of 8, 21, and 36 μg/ml for the 1, 10

105OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

and 25 mg/m3 groups, respectively. In females, th e post-exposure plasma levels were 1, 2, and 4 μg/ml for the 1, 10, and 25 mg/m3 groups, respectiv ely. When measured immediately before the daily exposure, plasma levels had re turned to baseline in females.

3.1.3.1.3 Dermal Exposure No specific dermal absorption studies have be en conducted in rats. However, Kennedy (1985) treated rats dermally with a to tal of 10 applications of APFO at doses of 0, 20, 200 or 2,000 mg/kg. Treatment resulted in elevated blood organofluorin e levels that increased in a dose-related manner. The results of in-vitro percutaneous absorption studi es of APFO through rat and human skin in vitro have been reported (Fasano et al, 2005). The pe rmeability coefficient for APFO was calculated to be 3.25 + 1.51 x 10 -5 cm/h and 9.49 + 2.86 x 10 -7 cm/h in rat and human skin, respectively.

3.1.3.2 Serum Pharmacokinetic Pa rameters in Adult Rats Serum pharmacokinetic parameters of PFOA have beenevaluated in adult Sprague-Dawley rats following gavage administration, and in Wist ar rats following i.v. administration.

3.1.3.2.1 Oral and Intravenous Exposure in Sprague-Dawley Rats Kemper (2003) examined the plasma concentrati on profile of PFOA following gavage and intravenous administration in sexually mature Sprague-Daw ley rats. Male and female rats (4/sex/group) were administered single doses of PFOA by gavage at dose rates of 0.1, 1, 5, and 25 mg/kg PFOA, and intravenously at a dose rate of 1 mg/kg PFO A. After dosing, plasma was collected for 22 days in males and 5 days in females. Plasma concen tration vs. time data were then analyzed by noncompartmental pharmacokinetic methods (Table 26 and Table 27). Comparison of AUC for the oral and intravenous 1 mg/kg doses in dicated that oral bioavailabil ity of PFOA was approximately 100%. Plasma elimination curves we re linear with respect to time inmale rats at all dose levels, while elimination kinetics were biphasic in fe males at the 5 mg/kg and 25 mg/kg dose levels. In males, plasma elimination half-lives were indepe ndent of dose level and ranged from approximately 138 hours to 202 hours. In females, terminal elimin ation half-lives ranged from approximately 2.8 hours at the lowest dose to approx imately 16 hours at the high dose. To further characterize plasma elimination kinetics, particularly in male rats, animals were given oral PFOA at a rate of 0.1 mg/kg, and plasma samples were collected until PFOA c oncentrations fell below quantitation limits (24 hours and 2016 hours in females and males, respectiv ely). Estimated plasma elimination half-lives in this experiment were approximately 277 hours in males and 3.4 hours in females (Table 26, Table 27).

106OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 26 Pharmacokinetic Parameters in Male Sprague-Dawley Rats Following Administration of PFOA by Gavage (Mean (SD))

Parameter 0.1 mg/kg 1 mg/kg 5 mg/kg 25 mg/kg 1 mg/kg
i.v. 0.1 mg/kg
extended time

T max(hr) 10.25 (6.45) 9.00 (3.83) 15.0 (10.5) 7.5 (6.2) NA 5.5 (7.0)
C max(μg/ml) 0.598 (0.127) 8.431 (1.161) 44.75 (6.14) 160.0 (12.0) NA 1.08 (0.42) Lambda z (1/hr) 0.004 (0.001) 0.005 (0.001) 0.0041 (0.0007) 0.0046 (0.0012) 0.004 (0.000) 0.0026 (0.0007)

T 1/2 (hr) 201.774 (37.489) 138.343 (31.972) 174.19 (28.92) 157.47 (38.39) 185.584 (19.558) 277.10 (56.62)
AUC INF (hr
. μg/ml) 123.224 (35.476) 1194.463 (215.578) 6733.70 (1392.83) 25,155.61 (7276.96) 1249.817 (113.167) 206.38 (59.03)
AUC INF /D (hr
. μg/ml/mg/kg) 1096.811 (310.491) 1176.009 (206.316) 1221.89 (250.28) 942.65 (284.67) 1123.384 (100.488) 2111.28 (586.77)
Cl p (ml/kg
. hr) 0.962 (0.240) 0.871 (0.158) 0.85 (0.21) 1.13 (0.31) 0.896 (0.082) 0.51 (0.17) (from Kemper, 2003)

Table 27 Pharmacokinetic Parameters in Female Sprague-Dawley Rats, Following Administration of PFOA by Gavage (Mean (SD))

Parameter 0.1 mg/kg 1 mg/kg 5 mg/kg 25 mg/kg 1 mg/kg i.v. 0.1 mg/kg extended time

T max (hr) 0.56 (0.31) 1.13 (0.63) 1.50 (0.58) 1.25 (0.87) NA 1.25 (0.50)

C max ( μg /ml) 0.67 (0.07) 4.782 (1.149) 20.36 (1.58) 132.6 (46.0) NA 0.52 (0.08) Lambda z (1/hr) 0.231 (0.066) 0.213 (0.053) 0.15 (0.02) 0.059 (0.037) 0.250 (0.047) 0.22 (0.07)

T 1/2 (hr) 3.206 (0.905) 3.457 (1.111) 4.60 (0.64) 16.22 (9.90) 2.844 (0.514) 3.44 (1.26)

AUC INF (hr
. μg /ml) 3.584 (0.666) 39.072 (10.172) 114.90 (11.23) 795.76 (187.51) 33.998 (7.601) 3.34 (0.32)

AUC INF /D (hr
. μg /ml/mg/kg) 31.721 (5.880) 38.635 (10.093) 20.78 (2.01) 29.54 (6.92) 30.747 (6.759) 34.39 (3.29)

Cl p (ml/kg
. hr) 32.359 (6.025) 27.286 (7.159) 48.48 (4.86) 35.06 (.88) 34.040 (9.230) 29.30 (3.06)
(from Kemper, 2003)

107OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

3.1.3.2.2 Intravenous Exposure in Wistar Rats Kudo et al. (2002) examined the plasma concentr ation profile of PFOA following i.v. administration in 9-week old Wistar rats. Male and female rats were injected intravenously with 48.63 μmol/kg body weight PFOA, and blood samples were collected at 5 min., 1, 6, 12 hrs, and 1, 2, 3, 4, 6, 8, 10 and 12 days after injection. A two-comp artment open model was used to estimate serum pharmacokinetic parameters (Table 28). The half-life of PFOA in male rats was found to be 70 times longer (5.7 days versus 1.9 hours) than in female s. In female rats, there appears to be biphasic elimination of PFOA; the fast phase occurs witha half-life of approximately 1.9 hours while the slow phase occurs with a half- life of approximately 24 hours. The total clearance of PFOA in female rats was 44 times higher than that in male rats.

Table 28 Pharmacokinetic Parameters of PFOA in Wistar Rats after i.v. Administration

Parameters Male Female Total clearance (ml/min per kg) 0.035 ± 0.010 1.551 ± 0.559*

T 1/2 (day) 5.68 ± 0.99 0.08 ± 0.03** Volume of distribution (ml/kg) 345.6 ± 57.3 211.2 ± 28.2* (from Kudo et al., 2002) Values are means ± S.D. for three to four rats. Differenc es were statistically significant between male and female rats (*, P <0.01: **, P<0.001).

3.1.3.3 Distribution Studies in Adult Rats

The distribution of PFOA has been examined in ti ssues of adult rats following administration by gavage and by i.v. and intraperitoneal (i.p.) inje ction. PFOA distributes pr imarily to the liver, plasma, and kidney, and to a lesser extent, other ti ssues of the body. It does not partition to the lipid fraction or adipose tissue. The distributi on of PFOA is predominantly extracellular.

3.1.3.3.1 Oral Exposure Kemper (2003) examined the distribution and clear ance of PFOA in tissues of male and female Sprague-Dawley rats following administration by gavage. Rats were administered 1, 5, and 25 mg/kg 14 C-PFOA by oral gavage. Tissue concentrations were determined at the time of maximum plasma concentration (Tmax) and at the time plas ma concentration had fallen to one half the maximum (T max/2 ). Values for T max and T max/2 for male and female rats were determined from pharmacokinetic experiments. T max/2 was calculated as T max + T 1/2 . For cases in which biphasic elimination was evident, the rapid phase T 1/2 was used for calculation of T max/2 . Tissues from male rats were collected at 10.5 hours (T max ) and 171 hours (T max/2 ) after dosing. Tissues from female rats were collected at 1.25 hours (T max ) and 4 hours (T max/2 ) after dosing. The results are summarized in Table 29 and Table 30 for males and females, respectively. Liver, kidney and blood were the primary tissues for distribution of 14C-PFOA. In males, the frac tion of the dose found in liver increased from T max to T max/2 , but remained constant or decreased in othe r tissues. In females, the fraction of the dose present in all tissues remained constant or decreased between T max and T max/2 . Liver-to-blood concentration ratios for 14 C at Tmax in males were greater than 1, and increased between T max and T max/2 . Kidney-to-blood concen tration ratios at T max in females were approxi mately 2 at all dose levels and remained relatively constant between T max and T max/2 .

108OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 29 Distribution of PFOA in Male Sprague-Dawley Rats after Oral Exposure (Percent of dose recovered at Tmax and Tmax/2 in tissues)

1 mg/kg 5 mg/kg 25 mg/kg Tissue T max T max/2 T max T max/2 T max T max/2 Prostate 0.083±0.039 0.030±0.002 0.071±0. 045 0.057±0.020 0.067±0.018 0.028±0.012
skin a 14.77±2.135 6.061±0.274 15.565±0.899 7.233±0.430 13.836±0.969 5.419±0.237
blood a 22.148±0.692 8.232±1.218 24.919±1.942 11.140±0.624 22.905±1.177 7.904±1.032
Brain 0.071±0.018 0.022±0.002 0.051±0.021 0.023±0.008 0.063±0.007 0.019±0.002
fat a 2.281±0.467 0.593±0.136 2.815±0.225 0.916±0.205 2.153±0.430 0.628±0.110
Heart 0.451±0.119 0.195±0.024 0.443±0.037 0.252±0.030 0.461±0.053 0.164±0.032 Lungs 0.74±0.147 0.341±0.043 0.593±0.376 0.344±0.194 0.863±0.103 0.303±0.057 Spleen 0.086±0.011 0.045±0.006 0.096±0. 017 0.060±0.007 0.106±0.015 0.042±0.005
Liver 21.708±5.627 32.627±3.601 18.750±2.434 25.231±1.289 17.528±0.900 20.145±3.098 Kidney 1.949±0.402 1.14±0.215 2.170±0. 354 1.212±0.115 2.293±0.286 1.003±0.122
G.I. tract 2.930±0.929 0.980±0.300 2.508± 0.713 1.052±0.202 2.784±0.608 0.808±0.189
G.I. contents 2.083±0.625 0.239±0.025 2.632±0.934 0.270±0.028 4.186±1.349 0.210±0.084

Thyroid 0.008±0.005 0.004±0.003 0.011±0. 006 0.004±0.002 0.009±0.002 0.005±0.001 Thymus 0.085±0.008 0.051±0.018 0.085±0. 012 0.053±0.003 0.120±0.025 0.045±0.010
Testes 0.755±0.079 0.356±0.037 0.693±0. 180 0.372±0.062 0.623±0.098 0.224±0.031
adrenals 0.019±0.004 0.010±0.001 0.022±0. 004 0.009±0.001 0.026±0.004 0.009±0.003
muscle a 12.025±0.648 4.984±0.745 13.565±0.576 6.429±0.648 12.855±0.841 4.253±0.358
bone a 3.273±0.538 1.120±0.094 3.047±0.544 1.375±0.169 3.062±0.438 0.906±0.100
Total b 85.465±6.426 57.026±3.379 88.033±1.420 56.031±1.025 83.937±3.680 42.112±4.740 (from Kemper, 2003) a Percent recovery scaled to whole anim al assuming the following: skin=19%, whole blood=7.4%, fat=7%, muscle=40.4%, bone=7.3% of body weight. b Totals are calculated from individual animal data.

Table 30 Distribution of PFOA in Female Sprague-Dawley Rats after Oral Exposure (Percent of dose recovered at Tmax and Tmax/2 in tissues)

1 mg/kg 5 mg/kg 25 mg/kg
Tissue T max T max/2 T max T max/2 T max T max/2 skin a 0.434±0.162 0.403±0.096 0.624±0.142 0.307±0.121 0.380±0.166 0.415±0.175
blood a 5.740±1.507 4.438±1.625 8.089±2.080 5.411±1.466 7.158±2.232 6.407±1.406
Brain 0.037±0.009 0.047±0.008 0.066±0.019 0.045±0.010 0.058±0.008 0.058±0.018
fat a 0.134±0.032 0.164±0.079 0.220±0.111 0.110±0.069 0.147±0.053 0.148±0.065
Heart 0.198±0.079 0.253±0.055 0.388±0.057 0.236±0.051 0.317±0.035 0.287±0.069 Lungs 0.454±0.148 0.546±0.082 0.827±0.102 0.570±0.179 0.678±0.067 0.775±0.204 Spleen 0.063±0.027 0.058±0.006 0.101±0. 021 0.060±0.012 0.091±0.007 0.070±0.002
Liver 7.060±1.266 6.817±1.537 11.190±2.1 92 7.176±0.982 10.538±1.723 9.080±0.895

109OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

110 1 mg/kg 5 mg/kg 25 mg/kg

Tissue T maxT max/2 T maxT max/2 T maxT max/2 Kidney 3.288±0.948 2.769±0.784 4.293±0. 771 2.685±0.736 5.867±0.946 4.749±0.393
G.I. tract 10.699±9.066 8.462±6.519 7.142± 2.594 8.255±8.967 6.923±1.846 3.547±1.306
G.I. contents 21.956±13.48 3.891±2.395 2.896±2.305 5.601±6.165 2.491±1.548 1.121±1.010

Thyroid 0.010±0.003 0.016±0.021 0.008±0. 002 0.006±0.002 0.009±0.003 0.007±0.002 Thymus 0.052±0.017 0.058±0.024 0.105±0. 030 0.068±0.021 0.091±0.032 0.077±0.020
Ovaries 0.047±0.019 0.048±0.006 0.071±0. 012 0.041±0.012 0.071±0.012 0.070±0.012
adrenals 0.014±0.005 0.018±0.004 0.026±0. 005 0.015±0.004 0.031±0.005 0.021±0.001
muscle a 0.170±0.051 0.258±0.089 0.325±0.010 0.229±0.031 0.441±0.116 0.304±0.099
Uterus 0.243±0.091 0.374±0.247 0.354±0. 046 0.247±0.068 0.358±0.124 0.365±0.029
bone a 0.101±0.017 0.153±0.052 0.174±0.057 0.142±0.078 0.157±0.072 0.181±0.090
Total b 50.698±16.48 28.772±10.98 36.897±3.187 31.201±12.63 35.803±2.554 27.680±2.569 (from Kemper, 2003) a Percent recovery scaled to whole animal assuming the following: skin=19%, whole blood=7.4%, fat=7%, muscle=40.4%, bone=7.3% of body weight. b Totals are calculated from individual animal data.

3.1.3.3.2 Intravenous Exposure Gibson and Johnson (1980) examined the tissue distribution in 10-week old male CD rats following a single i.v. dose of 13.1 mg/kg 14 C-APFO. The animals were sacrif iced 36 days after dosing. The results are presented in Table 31. PFOA distributed mainly to the liver, plasma and kidney.

Table 31 Distribution of PFOA in Ma le CD Rats After i.v. Exposure

Tissue Mean ± SD a
Liver 7.97 ± 4.02
Plasma 3.19 ± 1.72 Kidney 2.33 ± 1.13 Lung 0.79 ± 0.51
Red Blood Cells 0.77 ± 0.45
Skin 0.40 ± 0.28 Spleen 0.33 ± 0.19
Bone marrow 0.33 ± 0.26
Subcutaneous Fat 0.28 ± 0.18
Muscle 0.20 ± 0.09
Brain 0.11 ± 0.04
Abdominal Fat < 0.12 (from Gibson and Johnson, 1980) a Total 14C concentration expressed as μg equivalents of 14C-PFOA.OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

3.1.3.3.3 Intraperitoneal Exposure Ylinen et al. (1990) studied the distribution and e limination of PFOA in a limited sample of tissues following a single i.p. dose of 50 mg/kg in 10 week old Wistar ra ts (20/sex). Samples were collected from 2 animals per sex for analysis of PFOA at 12, 24 - 168 (in 24 hour intervals), 224 and 336 hours after dosing. The concentr ation of PFOA in the serum and tissues was determined with capillary gas chromatography equipped with a fl ame ionization detector (FID). A mass spectrometer was used in the selected ion monitoring mode when the PFOA concentration was below the quantitation limit of the FID (1 μg/ml). The half-l ife of PFOA was estimated from the linear regression of time and concentration of PFOA in a sem ilogarithmic plot. No PF OA was detected in the adipose tissue. In both sexes at 12 hours after administration, the highest concentration of PFOA was found in the serum, followed by the liver, kidne y, spleen and brain. In the females, the concentration of PFOA in the serum, liver and kidney decreased in a discontinuous fashion indicating distinct phases. The half-life of PFOA in the serum, liver, kidney and spleen is presented in Table 32; the values are generally much lower in females than in males.

Table 32 Mean Half-life (hours) of PFOA in Wistar Rats After i.p. Exposure

Tissue Males Females

Serum 105 24
Liver 210 60 Kidney 130 145 Spleen 170 73 (from Ylinen et al., 1990)

Vanden Heuvel et al. (1991) examined the distri bution of PFOA in the serum and a limited number of tissues following administration of 9.4 μmol/kg 14 C-PFOA by i.p. injection to 6-week old male and female Harlan Sprague-Dawley rats. At vari ous time-points for 28 days following treatment, four rats per sex were sacrificed; blood was co llected by cardiac puncture, and tissues were removed and frozen. 14 C-PFOA-derived radioactivity was quantitated using a Liquid Scintillation Analyzer. Assuming first order kinetics, tissu e elimination rates were calculated as T 1⁄2 = ln2/k e . The distribution and elimin ation of PFOA-derived 14 C in selected tissues is summarized in Table 33. In the male rats, 21% of the administ ered dose was present in the liver , and the next highest concentrations were found in the plasma and kidney. Far lowe r PFOA concentrations were found in the heart, testis, fat, and gastrocnemius muscle. In females, the highest concentratio ns of PFOA were found in the plasma followed by the kidney, liver and ovari es. In males, PFOA was eliminated from the liver at a slower rate than the other tissues; the T 1⁄2 for liver was 11 days compar ed to 8-9 days in most extrahepatic tissues. The rates of elimination were mu ch faster in the female rats than in the male rats. The high concentration of PFOA in the male live r was further examined using a liver perfusion technique. Liver was infused with 0.08 μmol 14 C-PFOA/min over a 48 min period for a total of 3.84 μmol 14 C. Approximately 11% of the cumulative dose of 14 C-PFOA infused was extracted by the liver in a first pass. At 2 min, the cumulative percent of PFOA extracted by the liver was 33%; that was substantially greater than the 11% cumulative dose of 14 C that was extracted after 48 min indicating that first-pass hepatic upta ke of PFOA may be saturable.

111OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 33 Distribution and Elimination of PF OA in Selected Tissues in Harlan Sprague-Dawley Rat after i.p. Exposurea

% Dose a Elimination Rate b Half-life
Tissue Male (2 hrs) Female (2 hrs) Male (24 hrs) Female (24 hrs) Male (day -1 ) Female (hour -1 ) Male (day) Female (hour)
Liver 2.03 (0.10) 1.53 (0.25) 2.08 (0.14) 0.06 (0.02) 0.062 (0.005) 0.185 (0.011) 11.3 3.8

Plasma 1.99 (0.03) 2.39 (0.44) 1.63 (0.11) 0.02 0.077 (0.006) 0.242 (0.007) 9.0 2.9

Kidney 0.95 (0.08) 2.00 (0.38) 0.74 (0.03) 0.06 (0.02) 0.074 (0.005) 0.214 (0.012) 9.4 3.2

Heart 0.42 (0.02) ND 0.39 (0.03) ND 0.066 (0.006) ND 10.4 ND

Gastrocnemius 0.26 (0.01) ND 0.21 (0.01) ND 0.076 (0.007) ND 9.2 ND

Fat c 0.32 (0.02) ND 0.27 (0.04) ND 0.087 (0.008) ND 8.0 ND

Testis/ Ovary 0.33 (0.01) 0.53 (0.10) 0.27 (0.03) 0.05 (0.01) 0.078 (0.006) ND 9.0 ND

(from Vanden Heuvel et al., 1991) ND Not determined a Values are percent dose of PFOA-deriv ed 14C per gram tissue; mean (SEM ); n=4; values with no SEM indicate that SEM < 0.01 b Mean (SEM); n=32 for males and n=16 for females c Epididymal fat pad

3.1.3.4 Metabolism Studies in Adult Rats

Several studies have examined metabolism of PF OA. However, no studies show clear evidence of metabolism. Ophaug and Singer (1980) found no change in ionic fluoride level in the serum or urine following oral administrati on of PFOA to female Holtzman ra ts. Ylinen et al. (1989) found no evidence of phase II metabolism of PFOA followi ng a single intraperitoneal PFOA dose (50 mg/kg) in male and female Wistar rats. Vanden Heuvel et al. (1991) investigated the meta bolism of PFOA in Harlan Sprague-Dawley rats administered 14 C-PFOA (9.4 μmol/kg, i.p.). Pooled daily ur ine samples (0-4 days post-treatment) and bile extracts analyzed by HPLC contained a single radioactive peak eluting identically to theparent compound. Tissues were taken from rats treated 4, 14, and 28 days after treatment to determine the presence of PFOA-containing lipid conjugates. Only the parent compound was present in rat tissues; no PFOA-containing hybri d lipids were detected. Fluoride concentrations in plasma and urine before and after PFOA treatment were un changed, indicating that PFOA does not undergo defluorination.

3.1.3.5 Elimination Studies in Adult Rats
In adult rats, there is evidence of enterohepatic circulation of PFOA. The urine is the major route of excretion of PFOA in the female rat, while the urin e and feces are both main routes of excretion in

112OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

male rats. There are gender differences in the elimination of PFOA in rats. 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. Hormonal changes during pregnancy do not appear to change th e rate of elimination in rats.

3.1.3.5.1 Enterohepatic Circulation Johnson et al. (1984) investigated th e effect of feeding cholestyramine to rats on the 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 of age, were give n a single i.v. injection of 13 mg/kg 14 C-APFO. 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, the mean per centage of PFOA eliminated in the feces of cholestyramine-treated rats was 9.8-fold the mean percenta ge eliminated in the feces of rats that did not receive cholestyramine. Excretion in urine was 41% of the administered dose for cholestyramine treated rats and 67% for rats that did not receive cholestyramine. 14 C in the liver equaled 4% or 12.1± 2.1 μg eq/g in cholestyramine treated rats and 8 % or 22.3 ± 6.2 μg eq/g in rats that did not receive cholestryamine. In plasma, the levels were 5.1 ± 1.7 μg eq/ml in choles tyramine treated rats and 14.7±6.8 μg eq/ml in rats that did not receive cholestyramine. In red blood cells, the levels were 1.8 ± 0.7 μg eq/ml in cholestryamine treated rats and 4.2 ± 2.4 μg eq/ml in rats that did receive cholestryamine. The high concentration of 14 C-APFO in the liver at 2 weeks after dosing and the fact that cholestyramine treatment enhances fecal elimination of 14 C nearly 10-fold suggests that there is enterohepatic circulation of PFOA.

3.1.3.5.2 General Elimination Studies

3.1.3.5.2.1 Oral Exposure Kemper (2003) investigated the el imination of PFOA in male and female Sprague-Dawley rats (4 rats/sex/group) administered a single dose of 14 C-PFOA by oral gavage at dose levels of 1, 5, and 25 mg/kg. Urine and feces were collected for 28 days in males and 7 days in females. Urine was the primary route of excretion of 14 C in both sexes, accounting for 43-62% of the administered dose in males and 76-84% of the administered dos e in females. Cumulative recovery of 14 C in the urine increased gradually over the 28 days inmale rats, but was essentially co mplete in female rats within the first 72 hours. Fecal excretion of 14 C accounted for 6-14% of the dose in males and 2-6% of the dose in females. Pilot experiments demonstrated that 14 C was not eliminated as either 14 CO 2 or volatile organic compounds in 14 C-PFOA-treated rats. Pretreatment of rats with 1 mg/kg-day PFOA for 14 days had little or no effect on the ex cretion of a challenge dose of 1 mg/kg 14 C-PFOA.

3.1.3.5.2.2 Intravenous Exposure Gibson and Johnson (1980) examined the excretion of total 14 C in male and female CD rats after a single i.v. dose of 14 C-APFO. The mean dose for females was 16.7 mg/kg while that for males was 13.1 mg/kg. Female rats excreted e ssentially all of the administered dose via the urine in the 24 hours after treatment. During the same time period, ma le rats excreted only 20% of the total dose. Male rats excreted 83% of the to tal dose via the urine and 5.4% vi a the feces by 36 days post dose. No radioactivity was detected in tissues of female rats at 17 days post dose ; 2.8% of the total dose was detected in the liver of male rats and 1.1% inthe plasma at 36 days post dose with lower levels equaling < 0.5% of the to tal dose in other organs.

113OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

3.1.3.5.3 Elimination Studies in the Pregnant Rat Hormonal changes during pregnancy do not appear to cause a change inthe rate of elimination of 14 C after oral administrati on of a single dose of 14 C-APFO (Gibson and Johnson, 1983). At 8 or 9 days after conception, four pregnant CD rats and two nonpregnant female CD rats were given a mean dose of 15 mg/kg 14 C-APFO. Individual urine samples were collected at 12, 24, 36, and 48 hours post dose and analyzed for 14 C content. Essentially all of the 14 C was eliminated via the urine within 24 hours for both groups of rats.

3.1.3.5.4 Studies on the Mechanism of the Gender Di fference in Elimination in Adult Rats

Several studies have been conducted to elucidate th e cause of the gender difference in rats in the elimination of PFOA. Hanhijarvi et al. (1982) conducted a series of st udies to examine the effect of probenecid, which inhibits the renal active secretion system for organic acids, on the elimination of PFOA in male and female Holtzman rats. In the firs t study, 4 male and 6 female Holtzman rats were administered 2 mg of nonionic fluo rine as PFOA by gavage. Seven fema le rats were administered 2ml distilled water as controls. The animals were then placed in metabolism cages and urine was collected until the animals were s acrificed at 24 hr by cardiac puncture. Serum was collected. Ionic fluoride and total fluorine content of serum and urine were determined, and nonionic fluorine was calculated as the difference. Twen ty-four hours after oral administration of PFOA, 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 μg/ml, 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 μg/ml. A mean of 97.5 ± 0.25% of the PFOA was bound in the plasma of both male and female rats. In the second study, Hanhijarvi et al. (1982) examined the effect of probenecid on the clearance of PFOA and inulin. Holtzman rats we re anesthetized and the femora l artery was cannulated for continuous infusion of 5% mannitol in isotonic saline , while the femoral artery was cannulated for drawing blood samples. The urinary bladder was also cannulated for serial collections of urine. When the urine and serum collections for the clearance study were complete, 65-68 mg/kg probenecid was administered by i.p. injection and after 20 - 30 minutes, additional 10 minute clearance tests were performed. Administratio n of probenecid reduced the PFOA /inulin clearance ratio in females from 14.5 to 0.46. PFOA clearance was re duced from 5.8 to 0.11 ml/min/100g. Net PFOA excretion was reduced from 4.6 μg/min/100g to 0.13 μg/min/100g. In male rats, however, thePFOA/inulin clearance ratio and the net excretion of PFOA were virtually unaffected by probenecid. In the males, PFOA clearance was 0.17 ml /min/100g, the PFOA/inulin clearance ratio was 0.22, and net PFOA excretion was 0.17 μg/min/mg. Finally, Hanhijarvi et al. (1982) examined the cu mulative excretion of PFO A over a 7-hour period. Holtzman rats were dosed i.v. with a mixture of 10%-20% radiolabel ed-PFOA and 80-90% unlabeled PFOA. Mannitol (5%) was infused and urine specimens were collected over 30-min intervals. The effect of probenecid was assessed by administ ering 65-68 mg/kg by i.p. injection at least 30 min prior to the administration of PFOA. Female rats excreted 76% of the administered dose of PFOA, while males excreted only7.8% of the administered dos e over a 7-hr period. Probenecid administration modified the cumulative excretion curve for males only slightly. However, in females probenecid markedly reduced PFOA eliminat ion to 11.8%. The authors concluded that the female rat possesses an active secretory mechan ism which rapidly eliminates PFOA from the body. Ylinen et al. (1989) studied the ur inary excretion of PFOA in male Wistar rats after castration and estradiol administration. Twenty male rats were castra ted at 28 days of age and were used in tests of PFOA excretion 5 weeks later. Ten castrated and 10 intact males were gi ven 500 μg/kg estradiol valerate by s.c. injection every second day for 14 days before administration of PFOA. PFOA was

114OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

administered as a single i.p. injection at 50 mg/ kg. Urine was collected inmetabolism cages for 96 hr after PFOA administration. Blo od samples were collected by card iac puncture. Six female rats were also included in the experime nt. 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% of the administered dose of PFOA, whereas the intact males excreted only 9 ± 4%. After the estradio l treatment, both the intact and cast rated males excreted PFOA in amounts similar to females, 61 ± 19% and 68 ± 14%, resp ectively. The castrated males without estradiol treatment excreted 50 ± 13% of th e administered dose of PFOA in th e urine. This was faster than the intact males but less than the females and the estrogen treated males. At the end of the test, the concentration of PFOA in the se rum of intact males was 17- 40 tim es higher than the concentration PFOA in the serum of other groups. There were no st atistically significant di fferences in the serum concentrations between the othe r groups. PFOA was similarly bound by the proteins in the serum of males and females. Vanden Heuvel et al. (1992a) investigated whethe r androgens or estrogens are involved in the marked sex-differences in the urinary excretion of PFOA. Castrated Harl an Sprague-Dawley male rats were given 9.4 μmol/kg 14 C-PFOA by i.p. injection. Castrati on increased the elimination of PFOA in the urine (36% of the dose was eliminated in 4 days versus 16% in controls), suggesting 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 e ffect on PFOA elimination whereas castration plus testosterone replacement at the physiological leve l reduced PFOA elimination to the same level as rats with intact testis. Thus, in male rats, testoste rone exerts an inhibitory effect on renal excretion of PFOA. In female rats, neither ovariectomy or ova riectomy plus testosterone affected the urinary excretion of PFOA, demonstrating th at the inhibitory effect of te stosterone on PFOA renal excretion is a male-specific response. Probenecid, which inhi bits the renal transport system, decreased the high rate of PFOA renal excretion incastrated males but had no effect on male rats with intact testis. Kudo et al. (2002) examined the role of sex hormones on the renal clearance (CL R ) of PFOA and the renal mRNA levels of specific organic anion transporters in male and female Wistar rats. Castration of male rats caused a 14-fold increase in CL R of PFOA. The elevated PFOA CL R in castrated males was reduced by treating them with testoster one. Treatment of male rats with estradiol increased the CL R of PFOA. In female rats, ovariectomy caused a significant increase in CL R of PFOA, which was reduced by estradiol treatment. Treatments of female rats with testosterone reduced the CL R of PFOA. Treatment with probenecid, a kno wn inhibitor of organic anion transporters, markedly reduced the CL R of PFOA in male rats, castrated ma le rats, and female rats. To identify the transporter molecules th at are responsible for PFOA tran sport 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 CL Rof PFOA. The level of organic anion transporter 2 (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 suggest ed that OAT2 and OAT3 are responsible for urinary elimination of PFOA in the rat; however, the po ssibility of a resorption process mediated by OATP1 was mentioned as a possible fact or in male rat retention of PFOA.

3.1.4 Metabolism and Pharmacokinetic Studies in Immature Rats

No studies have been conducted to specifically examine the abso rption, metabolism or elimination of PFOA in the developing rat. However, recent studies have b een conducted to examine the concentrations of PFOA in the developing Sprague-Daw ley rat, and to determine when the gender dif115OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

ference in elimination of PFOA becomes apparent. In addition, several studies have examined the serum and tissue distribution of PFOA in newly wean ed Wistar rats. These studies have shown that PFOA readily crosses the placenta and is present inthe breast milk of rats. During lactation and immediately after weaning, the elimination of PFOA is similar in males and females. In the male rat between 4-5 weeks of age, the factor(s) responsib le for the gender difference develop, and the rats assume the adult male elimination profile. In add ition, distribution studies in the postweaning rat have shown that PFOA is distributed prim arily to the serum, liver, and kidney.

3.1.4.1 PFOA Levels During Pregnancy and Lactation PFOA levels during gestation and lactation have b een studied (Hinderliter et al., 2005, Mylchreest, 2003). Pregnant Sprague-Dawley rats were dosed with 0, 3, 10 or 30 mg/kg-day APFO during days 4-10, 4-15, or 4-21 of ge station, or from gestation day 4 to lactation day 21. Clin ical observations and body weights were recorded daily. On gesta tion days 10, 15, and 21, 5 rats/group/time point were sacrificed and the number, location and type of implantation sites were recorded. Embryos were collected on day 10, and placent as, amniotic fluid, and embryos/f etuses were colle cted on days15 and 21. Maternal blood samples were collected at 2 hours ± 30 minutes post-dose. The remaining 5 rats per group were allowe d to deliver. On lactation days 0, 3, 7, 14, and 21, the pups were counted, weighed (sexes separate), and examined for abnormal appearance and behavior. Randomly selected pups were sacrificed and blood samples we re collected. On lacta tion days 3, 7, 14, and 21, the dams were anesthetized and milk and blood sa mples were collected; dams were removed from their litters 1-2 hours prior to co llection. Plasma, milk, amniotic fl uid extract, and tissue homogenates (placenta, embryo, and fetus) supernatants were analyzed for PFOA concentrations by HPLCMS.
All dams survived and there were no clinical si gns of toxicity. In the 30 mg/kg-day group, mean body weight gain was approximately 10% lower th an the control group during gestation, and mean body weights were approximately 4% lower than controls throughout gestation and lactation. The number of implantation sites, re sorptions, and live fetuses were comparable among groups on days 10, 15, and 21 of gestation. One dam in the 3 mg/kg-day group and two dams in the 30 mg/kg-day group delivered small litters (litte r size of 3-6 pups as compared to 12-19 pups/litter in the control group); however, given the small sample size the bi ological significance of th is finding is unclear. There were no clinical signs of toxicity in the pups, and pup survival and pup body weights were comparable among groups. Maternal PFOA levels during gestation and lacta tion are presented in Table 34. Maternal plasma levels at 2 hrs post-dosing (approximately the time of peak blood levels following a gavage dose) were fairly similar during the course of the study with a mean level of 11.2, 26.8, and 66.6 μg/ml in the 3, 10, and 30 mg/kg-day groups, respectively; PFO A levels in the control group were below the limit of quantitation (0.05 μg/ml). The concentration of PFOA in the milk was also fairly similar throughout lactation and was approximately 1/10th ofthe PFOA levels in the plasma; the mean values were 1.1, 2.8, and 6.2 μg/ml in the 3, 10, and 30 mg/kg-day groups, respectively.

116OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 34 Maternal PFOA Levels (μg/ml) During Gestation and Lactation a
Exposure Period Sample Time 3 mg/kg-day 10 mg/kg-day 30 mg/kg-day

GD 4 - GD 10 GD 10 plasma 8.53 ± 1.06 23.32 ± 2.15 70.49 ± 8.94
GD 4 - GD 15 GD 15 plasma 15.92 ± 12.96 29.40 ± 14.19 79.55 ± 3.11
GD 4 - GD 21 GD 21 plasma 14.04 ± 2.27 34.20 ± 6.68 76.36 ± 14.76
GD 4 - LD 3 LD 3 - plasma
- milk 11.01 ± 2.11
1.07 ± 0.26 22.47 ± 2.74
2.03 ± 0.33 54.39 ± 17.86
4.97 ± 1.20
GD 4 - LD 7 LD 7 - plasma
- milk 10.09 ± 2.90
0.94 ± 0.22 25.83 ± 2.07
2.74 ±0.91 66.91 ± 11.82
5.76 ±1.26
GD 4 - LD 14 LD 14 - plasma
- milk 9.69 ± 0.92
1.15 ± 0.06 23.79 ± 2,81
3.45 ± 1.18 54.65 ± 11.63
6.45 ± 1.38
GD 4 - LD 21 LD 21 - plasma
- milk 9.04 ± 1.01
1.13 ± 0.08 28.84 ± 5.15
3.07 ± 0.51 64.13 ± 1.45
7.48 ± 1.63
NA Average plasma Average milk 11.19 ± 2.76
1.07 ± 0.09 26.84 ± 4.21
2.82 ± 0.60 66.64 ± 9.80
6.16 ± 1.06 (from Hinderliter et al., 2005 and Mylchreest, 2003)

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

PFOA levels in the placenta, amniotic fluid, embr yo, fetus, and pup plasma are presented in Table 35. The levels of PFOA in the placenta on gestati on day 21 were approximately twice the levels observed on gestation day 15, and th e levels of PFOA in the amniotic fluid were approximately four times higher on day 21 than on day 15. The concentr ation of PFOA in the embryo/fetus was highest in the day 10 embryo and lowest in the day 15 embryo; PFOA levels in the day 21 fetus were intermediate. The concentration of PFOA in the plasmaof the day 21 fetus were approximately half the levels observed in the maternal plasma; the mean values were 5.9, 14.5, and 33.1 μg/ml in the 3, 10, and 30 mg/kg-day groups, respectively. Pup plasma le vels decreased until l actation day 7, and were thereafter similar to the levels observed in the milk.

Table 35 PFOA Concentrations (μg/ml) During Gestation and Lactation in Sprague-Dawley Rats a
Exposure Period Tissue 3 mg/kg-day 10 mg/kg-day 30 mg/kg-day
GD 4 - GD 10 GD 10 - embryo 1.40 ± 0.30 3.33 ± 0.81 12.49 ± 3.50
GD 4 - GD 15 GD 15 - placenta
- amniotic fluid - embryo 2.22 ± 1.79
0.60 ± 0.69
0.24 ± 0.19 5.10 ± 1.70
0.70 ± 0.15
0.53 ± 0.18 13.22 ± 1.03
1.70 ± 0.91
1.24 ± 0.22
GD 4 - GD 21 GD 21 - placenta
- amniotic fluid
- fetus - fetal plasma 3.55 ± 0.57
1.50 ± 0.32
1.27 ± 0.26
5.88 ± 0.69 9.37 ± 1.76
3.76 ± 0.81
2.61 ± 0.37
14.48 ± 1.51 24.37 ± 4.13
8.13 ± 0.86
8.77 ± 2.36
33.11 ± 4.64
GD 4 - LD 3 LD 3 - pup plasma 2.89 ± 0.70 5.94 ± 1.44 11.96 ± 1.66
GD 4 - LD 7 LD 7 - pup plasma 0.65 ± 0.20 2.77 ± 0.58 4.92 ± 1.28
GD 4 - LD 14 LD 14 - pup plasma 0.77 ± 0.10 2.22 ± 0.38 4.91 ± 1.12

117OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

118

Exposure Period Tissue 3 mg/kg-day 10 mg/kg-day 30 mg/kg-day
GD 4 - LD 21 LD 21 - pup plasma 1.28 ± 0.72 3.25 ± 0.52 7.36 ± 2.17 (from Hinderliter et al., 2005 and Mylchreest, 2003) a mean ± SD; samples were pooled by litter and were collected 2 hrs post-dosing

3.1.4.2 PFOA Levels in the Postweaning Sprague-Dawley Rat

Two studies have been conducted to examine PF OA levels in the postweaning Sprague-Dawley rat to determine the ontogeny of the gender difference in elimination. Han (2003) administered four to eight week old rats (10/sex/age) a single dose of 10 mg/kg-day APFO by oral gavage. Blood samples were collected 24 hours af ter dosing and the plasma concen tration of PFOA was measured by HPLC-MS. At four weeks of ag e, the concentration of plasmaPFOA was approximately 2.7 times higher in males than in the females (Table 36). In females, the concentration of plasma PFOA decreased by 2.7 fold between 4 and 5 weeks of age, and thereafter remained fairly steady. In males, the concentration of plasma PFOA increased by 5.4 fold between 4 and 5 weeks of age, and thereafter remained fairly steady. Between 5 and 8 weeks of age, the PFOA plasma concentrations were 34.7 - 65.1 fold higher in males than in females of th e same age. Thus, it appears that maturation of the factor(s) responsible for the gender difference in elimination of PFOA occurs between the ages of 4 and 5 weeks in the rat.

Table 36 Plasma PFOA Concentrations (μg/ml) in Postweaning Sprague-Dawley Rats a
Age (weeks) Males Females

4 7.32 ± 1.01 2.68 ± 0.64
5 39.24 ± 3.89 1.13 ± 0.46
6 43.19 ± 3.79 1.18 ± 0.52
7 37.12 ± 4.07 0.57 ± 0.29
8 38.55 ± 5.44 0.81 ± 0.27 (from Han, 2003)

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

Hinderliter (2004) investigated th e relationship between age and both plasma and urine PFOA concentrations in male and female Sprague-Dawley rats . Immature rats 3, 4, or 5 weeks of age were administered APFO via oral gavage as a single dose of 10 or 30 mg/kg. Rats were not fasted prior to dosing. Two hours after dosing, 5 rats /sex/group were sacrificed and blood samples were collected. The remaining 5 rats/sex/group were placed in metabolism cages for 24-hour urine collection. These rats were sacrificed at 24 hours and blood samples were co llected. In male rats, plasma PFOA concentrations for either the 10 or 30 mg /kg dosage groups did notdiffer significantly by sample time (at 2 and 24 hours) or by animal age ( 3, 4 or 5 weeks), except at 2 hours and five weeks of age (p<0.01) which showed the lowest PFOA le vel (Table 37). PFOA plasma concentrations following a 30 mg/kg dose were 2-3 times highe r than those following a 10 mg/kg dose.OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 37 Plasma PFOA Concentrations in Male Rats Dosed via Oral Gavage
Plasma PFOA ( μ g/ml)

2 Hours Post-Dose 24 Hours Post-Dose Age (weeks) Dose (mg/kg)

Mean SD Mean SD
3 10 41.87 4.01 34.22 7.89
4 10 39.92 4.45 42.94 5.33
5 10 26.32* 6.89 40.60 3.69
3 30 120.65 12.78 74.16 18.23
4 30 117.40 18.10 100.81 13.18
5 30 65.66* 15.53 113.86 23.36 (from Hinderliter 2004) *Statistically significantly different by sample time and animal age (p<0.01). In female rats, plasma PFOA concentrations were si gnificantly lower at 5 weeks of age than at 3 or 4 weeks of age for both sample times and doses, except at 2 hours post-dose in the 10 μ g/kg dose group (p<0.01, Table 38). Plasma PFOA concentra tions were significantly lower at 24 hours post dosing as compared to 2 hours pos t dosing in all ages and doses tested. Plasma PFOA concentrations following a 30 mg/kg dose were approximatel y 1.5 to 4 times higher than those observed following a 10 mg/kg dose. At 24 hours post dose, plasma PFOA levels in the female rats was significantly lower than the plasma PFOA levels in male rats, especially at 5 weeks of age.

Table 38 Plasma PFOA Concentrations in Female Rats Dosed via Oral Gavage

Plasma PFOA ( μ g/ml)

2 Hours Post-Dose 24 Hours Post-Dose Age (weeks) Dose (mg/kg)

Mean SD Mean SD
3 10 37.87 5.77 13.55 b 3.83
4 10 29.88 12.15 18.98 b 7.01
5 10 33.23 7.41 1.36 a, b 0.87
3 30 84.86 10.51 51.43 b 13.61
4 30 80.67 14.10 28.01 b 9.90
5 30 56.90 a 29.66 3.42 a, b 1.95 (from Hinderliter, 2004) a Statistically significantly different from the 3 and 4 week values (p<0.01). b Statistically significantly different from 2 hour values (p<0.01). The results of the prior study (Han, 2003, Table 15) are consistent w ith the current study for plasma PFOA concentrations of both sexes at 5 weeks of ag e but not at 4 weeks of age (3 week old animals were not previously tested) when rats were admi nistered 10 mg/kg APFO and plasma was collected at 24 hours post-dose. The only known variation betw een the two studies is that in the previous study, rats were fasted overnight for approximat ely 12 hours before dosing with APFO. In a supplemental study to determine the effect of fasti ng (Hinderliter, 2004), 4 week old rats, 4 rats/sex, were administered 10 mg/kg APFO via oral gavage. Two animals/sex were fasted overnight for 12

119OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

hours before dosing with APFO. All the rats were sacrificed at 24 hours post dosing and blood was collected for analysis of PFOA in plasma. Plasma PFOA concentrations in male rats were 64.95 and 30.00 μ g/ml for the fasted and non-fasted animals, re spectively. Plasma PFOA concentrations in the female rats were 68.16 and 26.54 μ g/ml for the fasted and non-fast ed animals, respectively. Given the consistency in the 5 week ol d rat plasma PFOA concentrations, the authors concluded there may be an unknown physiological variability in rats younger than 5 weeks which impacts the pharmacokinetics of PFOA.
Urine PFOA concentrations differe d significantly with age, dose an d sex (p<0.01, Table 39). Female rats had higher urine PFOA concentrations th an males, and the female urine PFOA concentrations increased with age. In male rats, urine PF OA concentrations decreased with age. In both sexes, urine PFOA was higher (2.5 to 6.5 times) at the 30 mg/kg dose as compared to the 10 mg/kg dose. No attempt was made to quantify urinary out put (volume) or standardize for creatinine levels.

Table 39 Urine PFOA Concentrations in Male and Female Rats Dosed via Oral Gavage

Urine PFOA ( μ g/ml at 24 hours post-dose)
Male Female Age (weeks) Dose (mg/kg)

Mean SD Mean SD
3 10 9.57 4.86 21.17 8.95
4 10 4.53 2.45 23.26 15.27
5 10 4.03 2.36 49.77 24.64
3 30 51.76 28.86 94.89 26.26
4 30 28.70 18.84 104.12 28.97
5 30 15.65 6.24 123.16 51.56 (from Hinderliter 2004)

3.1.4.3 Serum and Tissue Distribution in Immature Wistar Rats Following Oral Exposure Ylinen et al. (1990) administered newly weaned Wistar rats (18/ sex/group) doses of 3, 10, and 30 mg/kg-day PFOA by gavage for 28 days. At necr opsy, serum was collected as well as the brain, liver, kidney, lung, spleen, ovary, te stis, and adipose tissue. The c oncentration of PFOA in the serum and tissues was determined with capillary gas chromatography equipped with a flame ionization detector (FID). A mass spectrometer was used in the selected ion monitoring mode when the PFOA concentration was below the quantitation limit of the FID (1 μg/ml). The concentration of PFOA in the serum and tissues following 28 days of administration is pres ented in Table 40. PFOA was not detected in the adipose tissue. The concentrations of PFOA in the serum and tissues were much higher in males than in females. In the males, the levels of PFOA in the serum and tissues were generally lower in the 30 mg/kg-day group than in the 10 mg/kg-day group du e to increased urinary elimination inthe 30 mg/kg-day group.

120OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 40 Tissue Distribution of PFOA in Wistar Rats after 28 Days of Treatment

Males a Females a

Tissue 3 mg/kg-day 10 mg/kg-day 30 mg/kg-day 3 mg/kg-day 10 mg/kg-day 30 mg/kgday

Serum 48.6±10.3 87.27±20.09 51.65±1.47 2.4 b 12.47±4.07 13.92±6.06
Liver 39.9±7.25 51.71±11.18 49.77±10.76 1.81±0.49 3.45±1.36 6.64±2.64 Kidney 1.55±0.71 40.56±14.94 39.81±17.67 0.06±0.02 7.36±3.19 12.54±8.24 Spleen 4.75±1.66 7.59±3.5 4.1±1.57 0.15±0.04 0.38±0.17 1.59±0.49 Lung 2.95±0.54 22.58±4.59 23.71±5.42 0.24 b 0.22±0.15 0.75±0.26
Brain 0.398±0.144 1.464±0.211 0.71±0.32 < LOQ c 0.029±0.019 0.044±0.018 Ovary < LOQ 0.41±0.27 1.16±0.58
Testis 6.24±2.04 9.35±4.02 7.22±3.17 (fromYlinen et al., 1990) a N = 6, Mean ± SD, μg/ml b N=3, no SD c below the limit of quantitation Hanhijarvi et al. (1987) admi nistered PFOA by gavage to 48 newly-weaned Wistar rats (6/sex/group) at 0, 3, 10, and 30 mg/kg-day for 28 consecutive days and determined the serum levels of PFOA. At the end of the study, blood was co llected via cardiac puncture; PFOA levels were determined by gas chromatography. 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 μg/ml, 83.1 ± 24.7 μg/ml, and 53.4 ± 11.2 μg/ml, respectively for the 3, 10 and 30 mg/kg-day dose levels. The corresponding figures for female rats were 2.43 ± 5.96 μg/ml, 11.3 ± 8.59 μg/m l, and 9.06 ± 8.80 μg/ml respectively, for the 3, 10 and 30 mg/kg-day dose levels.

3.1.5 Comparative Studies of Protein Binding in Humans, Non-Human Primates, and Rats It has been suggested that PFOA circulates inthe body by noncovalently binding to plasma proteins. Several studies have inves tigated the binding of PFOA to plasma proteins of rats, humans or monkeys to gain understanding of its absorption, distribution and e limination, and species and gender differences. Protein binding in plasma from cynomolgus monkeys, rats, and humans was tested with PFOA (Kerstner-Wood et al., 2003). The re sults are summarized in Tabl e 41. Rat, human, and monkey plasma all bind PFOA at 97- 100% at tested concentrations ranging from 1-500 ppm. PFOA was found to bind primarily to human serum albumin co mpared to other protei n components of human plasma at physiological concentrations.

121OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 41 % Protein Binding in Rat, Human, and Monkey Plasma

PFOA Concentration (ppm) Rat Monkey Human

1 ~100 ~100 ~100
10 99.5 99.8 99.9
100 98.6 99.8 99.9
250 97.6 99.8 99.6
500 97.3 99.5 99.4 (from Kerstner-Wood et al., 2003) % Binding values reported as “~100" reflect a nonquantifia ble amount of test article in the plasma water BQL<6.25 ng/ml Han et al. (2003) investigated the binding of PFOA to rat and human plasma proteins in vitro. Male and female rats treated in vivo showed no gend er difference in the binding of PFOA to serum, though the persistence of PFOA in vi vo is much greater in male thanfemale rats. The authors conclude that there is no correlation between the PF OA persistence and binding of the PFOA to rat serum. The primary PFOA binding protein in plasmawas serum albumin. However, the method used (ligand blotting) would not theore tically allow the iden tification of low abunda nce proteins with high affinity for PFOA. Further investigation of purified rodent and human serum albumin binding using labeled 19 F NMR allowed the calculation of disassociation constants for PFOA binding to rodent and human serum albumin. No significant differences between bi nding to the two proteins was detected (Table 42).

Table 42 Dissociation Constants ( K d

) of Binding between PFOA and Rodent Serum Albumin (RSA) and Human Serum Albumin (HSA) and the Number of PFOA Binding Sites ( n ) on RSA and HSA

Parameter Method RSA HSA

K

d (mM) NMR a 0.29 ± 0.10 c

K

d (mM) micro-SEC b 0.36 ± 0.08 c 0.38 ± 0.04
N micro-SEC b 7.8 ± 1.5 7.2 ± 1.3 (from Han et al., 2003) a Average of the two Kd values (0.31 ± 0.15 and 0.27 ± 0.05 mM) obtained by NMR. b Values were obtained from three independent experi ments and their standard deviations are shown. c On the basis of the result of unpaired t-test at 95% c onfidence interval, the difference of Kd values determined by NMR and micro-SEC is statistically insignificant.

3.1.6 Metabolism and Pharmacokinetic Studies in Other Test Species

There is limited information on the metabolism and pharmacokinetics of PFOA in mice, rabbits and dogs. No specific pharmacokinetic studies have beenconducted in mice. However, toxicology studies in mice indicate that PFOA is absorbed, and furthermore, there does not appear to be a gender difference in elimination. For example, Sohlenius et al. (1992) exposed male and female C57B1/6 mice to dietary levels of 0.02% PFOA for one wee k. There was a significant decrease in mean body weight and a significant increase in absolute liver weight; the res ponse was similar in male and female mice. In rabbits, there is no information available on th e metabolism or eliminati on of PFOA by any route of exposure. No specific studies of the absorptio n of PFOA have been c onducted following oral or

122OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

inhalation exposure. However, there is evidence that PFOA is absorbed following dermal exposure. O’Malley and Ebbens (1981) treatedmale and female New Zealand White rabbits dermally with doses of 100, 1,000 and 2,000 mg/kg APFO for 14 days . Mortality was 100% (4/4) in the 2,000 mg/kg group, 75% (3/4) in the 1,000 mg/kg group a nd 0% (0/4) in the 100 mg/kg group. Similarly, Kennedy (1985) treated rabbits dermally with a total of 10 applications of APFO at doses of 0, 20, 200 or 2,000 mg/kg. Treatment resulted in elevated bl ood organofluorine levels that increased in a dose-related manner. In addition, limited information is available on the serum and liver distribution of PFOA in rabbits following i.v. administration. Johnson (1995a) administ ered individual female rabbits intravenous doses of 0, 4, 16, 24 and 40 mg/kg te trabutyl ammonium salt of PFO A. The animal given 40 mg/kg died within 5 minutes of trea tment. All other animals appear ed normal throughout the study. Serum samples were analyzed for total organic fluorin e at 2, 4, 6, 8, 12, 24, and 48 hours post dose. At 2 hrs, serum organic fluorine levels in the rabb its that received 0, 4, 16, and 24 mg/kg were 1.25 μg/ml, 4.09 μg/ml, 14.9 μg/ml, and 41.0 μg/ml, re spectively. There was a rapid decrease of total organic fluorine in the serum with time; it wa s essentially eliminated at 24 hr and was nondetectable at 48 hr. The biological half-life was on the order of 4 hours. The total organic fluorine levels in whole liver at 48 hr post dose for the rabbits that received 0 mg/kg, 4 mg/kg, 16 mg/kg, and 24 mg/kg were 20 μg, 43 μg, 66 μg, and 54 μg, respectively. There is no information on the sp ecific absorption, metabolism or di stribution of PFOA in dogs by any route of exposure. One study examined the e limination of PFOA in dogs following i.v. administration. Hanhijarvi et al. (1988) ad ministered beagle dogs (3/sex) an i.v. injection of 30 mg/kg of PFOA followed by continuous infusion with 5% mannitol. Urine and blood were collected at 10 minute intervals for 60 min. Probenecid was then administered by i.v. injection, and urine and blood samples were collected as before. Renal clearanc e of PFOA was calculated for the before and after probenecid injection periods. Four additional dogs (2/sex) were given 30 mg/kg of PFOA by i.v. injection. These dogs were kept in metabolism cages, and blood samples were collected intermittently for 30 days. The renal clearance rate wa s approximately 0.03 ml/min/kg. Probenecid significantly reduced the PFOA clearance rate in both sexe s, indicating an active secretion mechanism for PFOA. The plasma half-life of PFOA was 473 hr be fore probenecid administration and 541 hr after in male dogs, and 202 hr before probeneci d and 305 hr after in the female dogs.

Conclusions Little information is available concerning the pharmacokinetics of PFOA and its salts in humans. Preliminary results of a 5-year half-life study in 26 retired worker s indicate that the mean serum elimination half-life of PFOA in these workers was 3.8 years (1378 days, 95% CI, 1131-1624 days) and the range was 1.5 - 9.1 years. Metabolism and pharmacokinetic studies in non-human primates has been examined in a study of 3 male and 3 fe male cynomolgus monkeys administered a single i.v. dose of 10 mg/kg potassium PFOA. In male monkeys, the av erage serum half-life was 20.9 days. In female monkeys, the average serum ha lf-life was 32.6 days. In addition, 4-6 male cynomolgus monkeys were administered APFO daily via oral capsule at 10 or 20 mg/kg-day for six months, and the elimination of PFOA was mon itored after cessation of dosing. For the two 10 mg/kg-day recovery monkeys, serum PFOA elim ination half-life was 19.5 days, and the serum PFOA elimination half-life was 20.8 days for the three 20 mg/kg-day monkeys. Studies in adult rats have shown that the ammonium salt of PFOA (APFO) is absorbed following oral, inhalation and dermal exposure. Serum pharmacokinetic parameters and the distribution of PFOA has been examined in the tissues of adult rats following administration by gavage and by i.v. and i.p. injection. PFOA distributes primarily to the liver, serum, and kidney, and to a lesser extent, other tissues of the body. It does not partition to the li pid fraction or adipos e tissue. PFOA is not metabolized and there is eviden ce of enterohepatic ci rculation of the compound. The urine is the

123OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

major route of excretion of PFOA inthe female rat, while the urine and feces are both main routes of excretion in male rats. There are gender differences in the elimination of PFOA in adult rats following administration by gavage and by i.v. and i.p.injection. In female rats , following oral administration, estimates of the serum half-life were dependent on dose and ranged from approximate ly 2.8 to 16 hours, while in male rats estimates of the serum half-life following oral administration were independent of dose and ranged from approximately 138 to 202 hours. In fe male rats, elimination of PFOA appears to be biphasic with a fast phase and a sl ow phase. The rapid excretion of PFOA by female rats is believed to be due to active renal tubular se cretion (organic anion tr ansporters); this rena l tubular secretion is believed to be hormonally controlled. Hormonal ch anges during pregnancy do not appear to cause a change in the rate of elimination in rats.
Several recent studies have beenconducted to examine the kinetics of PFOA in the developing Sprague-Dawley rat. These studies have shown that PFOA readily cr osses the placenta and is present in the breast milk of rats. The gender difference in eliminati on is developmentally regulated; between 4-5 weeks of age, elimination assumes th e adult pattern and the gender difference becomes readily apparent. Distribution studi es in the postweaning rat have shown that PFOA is distributed primarily to the serum, liver, and kidney.

3.2 Epidemiology Studies 3M and Dupont have conducted several epidemio logy and medical surveillance studies of the workers at their plants in various cities of the U.S. No remarkable health effects that can be directly attributed to PFOA exposure have been reported in fluorochemical producti on workers described in the studies below.

3.2.1 Mortality and Cancer Incidence Studies in Workers 3M and Dupont have both conducted mortality and cancer incidence studies in workers exposed to POSF-based fluorochemicals. Mortality studies have been undertaken at 3M’s Cottage Grove plant in Minnesota while Dupont carried out mortality and cancer incidence studies on workers employed at their Washington Works pl ant in West Virginia. A retrospective cohort mortality study was performed on employees at the 3M 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 de ath. There was almost complete follow-up (99.5%) of all of the study participants. The exposure stat us 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 Divi sion provided the cumulative expos ure measurements. Of the 3537 (2788 men and 749 women) employees who partic ipated 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. Standardized Mortality Ratios (SMR s), 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 stratifie d for 3 latency period s (10, 15, and 20 years) and 3 periods of duration of employment (5, 10, and 20 years).

124OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

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 an d 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 th e causes of death when compared to both the U.S. and the Minnesota d eath rates. When latency and duration of employment were considered, there were no elevated SM Rs. When employee deaths in the Chemical Division were compared to Minnesota death rates, the SMR for pros tate 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 significan t (p = 0.03) association with length of employment in the Chemical Division and prostate cancer mortality. Based on the results of proportional ha zard models, the relative risk for a 1-year in crease 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 c ohort consisted of 3992 workers. Th e 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, j obs 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 fluoroche micals (n = 1815, primarily non-chemical 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-expos ed 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 we re 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 fr om pancreatic cancer (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 experience cerebrovascular disease mortal ity (5 deaths observed, 1.94 expected; 95% CI = 0.84 – 6.03). The excess mortality in cerebrovasc ular disease noted in employees in the definite exposure group was further analyzed based on numbe r of years of employment at the plant. Three of the 5 deaths occurred in workers who were employed in jobs with definite PF OA 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 expos ure. In order to confirm that the results regarding cerebrovascular disease were not an artifact of death certificat e 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 acco rding 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 cerebrovasc ular disease than the general population. A doseresponse relationship was not obser ved with years of exposure.

125OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

In the probable exposure group, 3 SMRs were elevated: cancer of the testis and other male genital organs (SMR = 2.75, 95% CI = 0.07 – 15.3); panc reatic 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 pros tate 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 pros tate 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 gr oup. All of these deaths would have been placed in the chemical plant employee 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 pros tate 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 coul d be altered significantly. This issue has become more apparent, given the results of a biomonitori ng study that took place at the Cottage Grove plant in 2000 in which PFOA concentrations were not corr elated with years worked in the Chemical Division but instead were associated with the specific area of the plant where APFO was produced (Olsen, et al., 2003f). The exposure categories were modified between th e first and second Cottage Grove mortality studies; therefore, comparison of the results is difficult. 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 exposur e categories that were de fined without measured exposures. Although serum PFOA measurements were considered in the expos ure matrix developed for the update, they were not directly used. In th e second study, the 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 needs to continue. Although there were more than 200 additional deaths included in this analysis, it is a small number and the cohort is s till relatively young. Given the results of studies on fluorochemicals in both animals and humans, further analysis is warranted. Limited data are available on mortality and cancer incidence at Dupont’s Washington Works Plant in Parkersburg, WV. These studies were periodically undertaken as part of a medical surveillance program at the plant. The most recent report is summarized here. Cancer incidence for active employees was reported for the 1959-2001 time period and mortality data were reported for active and retired employees for 1957 thr ough 2000 (Dupont, 2003). No other data, such as employee exposure information, lifestyle factors, employee demogra phics, or other chemicals used at the plant, are available in this report. In the cancer incidence report, cancer cases we re identified through a co mbination of company health and life insurance claims and company can cer and mortality registri es. Standardized incidence ratios (SIRs) were only calculated for those cancers for which 5 or more cases were observed, which included 14 types of cancer. Two of those cancer types were elevated and statistically significant (p = 0.05): bladder [SIR = 1.9; 95% CI (1.15-3.07)] and kidney and ur inary organs [SIR =2.3 (95% CI = 1.36-3.65)]. All of the reported cases were male. Some other cancer types with ele126OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

vated SIRs but which were not statistically significant at p = 0.05 included myeloid leukemia (2.02), cancer of the larynx (1.77), multiple myel oma and immunoproliferative (1.72), malignant melanoma of skin (1.3), testicular can cer (1.46), and brai n cancer (1.2). In the mortality report, when all causes of death were reported, SMRs, adjusted for age and gender, were statistically significant (p = 0.05) for rheumatic heart disease (SMR = 3.55; 95% CI, 1.14- 8.30) and atherosclerosis and aneury sm (SMR = 1.98; 95% CI, 1.17-3.14). Two separate analyses of leukemia incidence were conducted prior to the a bove studies at Dupont’s Washington Works plant (Walrath and Burke, 1989; Karns and Fayerweather, 1991). The initial study reported a statistically significant (p = 0.10) elevated odds ratio (OR) of 2.1 for leukemia incidence for male employees working at the plan t from 1956-1989. Eight cases (all male) of different types of leukemia were identified. The OR remained elevated when the workers were divided into wage and salaried employees (2.2 and 2.0, respectiv ely). In the follow-up case-control study, four controls were selected from the plant for each case, matched on gender, age and payroll status. Matched odds ratios were significantly elevated (p = 0.10) for employees who had previously worked as custodians and engineers, 8.0 (90% CI, 1.1-60.0) and 7.9 (90% CI, 1.0 - 76.0), respectively and remained elevated (although not statisti cally significant) for these same job categories within the plant (OR= 4.0 and 5.1, respectively). Matched OR were also reported based on the area of the plant where the cases work ed; however, no statistic ally significant (p = 0.10) OR were reported. The mortality data reported above do not show any statistically signifi cant (p = 0.10) elevations in leukemia deaths (all of the cases in the case-contro l study were dead at the time of the mortality report), possibly because the number of cases was very small and divided among different types of leukemias. The Washington Works data provide some insight as to where more medical surveillance should be concentrated at this plant but pr ovide little information about the relationship of PFOA to mortality or cancer incidence since no exposure information, use of other chemicals, or lifestyle information was collected on these employees.

3.2.2 Hormone Study in Male Workers

Endocrine effects have been associated with PFOA exposure in animals; therefore, medical surveillance data, including hormone testing, from male employees only of the Co ttage Grove, Minnesota plant were analyzed (Olsen et al., 1998a). PFOA serum leve ls 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 and 0-115 ppm in 1995 using thermospray mass spectrophotometry assay. Eleven hormones were assayed from the serum samples. They were: cortisol, dehydroepiandrosterone sulfat e (DHEAS), estradiol, follicle stimulating hormone (FSH), 17 gamma-hydroxyprogesterone (17-HP), free testoster one, total testosterone, luteinizing hormone (LH), prolactin, thyroid-stimul ating 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 an alysis, ANOVA, and Pearson correlation coefficients. PFOA was not highly correlated with any of the hormones or with the following covariates: age, alcohol consumption, body mass index (BMI), or ci garettes. Most of the employees had PFOA serum levels less than 10 ppm. In 1993, only 12 employees had se rum 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. Therefor e, it is likely that there was not enough power to detect differences in either of the highest categories. The mean age

127OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

of the employees in the highest exposure category was the lowestin both 1993 and 1995 (33.3 years and 38.2 years, respectively). Although not significan tly different from the other categories, BMI was slightly higher in th e highest PFOA category.
When the mean values of the various hormones we re compared by exposure categories, there was a statistically significant (p = .01) elevation in prolactin in 1993 onl y for the 10 workers whose serum levels were between 10 and 30 ppm compared to th e lower 2 exposure categories. In addition, TSH was significantly (p = .002) elevated in the same exposure category for 1995 only (mean blood serum level was 2.9 ppm). However, mean TSH levels for the other exposure categories, including the ≥ 30 ppm category, were all the same (1.7 ppm). In 1993, TSH was elevated only in this same exposure category, as well; however did not re ach statistical sign ificance (p = .09). Estradiol levels in the >30 ppm group in both years were 10% higher than the other PFOA groups; however, the difference was not st atistically significant (p <0.05) . These results were confounded by estradiol being 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. 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 studi es (55 ppm PFOA in the CD rat). Only 3 employees in this study had PFOA serum leve ls this high. They also suggest that the higher estradiol levels in the highest exposure category could suggest a threshold relationship between PFOA and estradiol. The authors did not report a negative association between PFOA serum le vels and testosterone. There were no statistically signi ficant trends (p < 0.05) noted for PFOA and either bound or free testosterone. However, 17-HP, a precursor of testosterone, was hi ghest in the >30 ppm PFOA group in both 1993 and 1995. In 1995, PFOA wa s significantly associated wi th 17-HP in regression models adjusted for possible confounders. However, the authors state that this a ssociation was based on the results of one employee (data were not provide d in the report). Free test osterone was highly correlated with age in both 1993 and 1995 (r = -.48, p < .001; r = -.40, p < .001, respectively). There are several design issues that should be noted when evaluating the results of this study. First, although there were 2 study years (1993 and 1995), th e populations were notindependent. Sixtyeight employees participated in both years. Se cond, there were 31 fewer employees who participated in the study in 1995, thus reducing the power of the study. Th ere were also very few employees in either year with serum PFOA levels greate r than 10 ppm. Third, the cross-sectional design of the study does not allow for analysis of temporality of an association. Since the half-life of PFOA is at least 1 year, the authors suggest that it is pos sible that there may be some biological accommodation to the effects of PFOA. F ourth, only one sample was taken for each hormone for each of the study years. In order to get more accurate measurements for some of the hormones, pooled blood taken in a short time period should have been used for each particip ant. Fifth, some of the associations that were measured in this study were done based on the results of an earlier paper that linked PFOA with increased estradiol and decreased test osterone levels. Howeve r, total serum organic fluorine was measured in that study instead of PFOA, making it difficult to compare the results. Finally, there may have been some measurement e rror of some of the confounding variables.

3.2.3 Occupational Study on Episodes of Care In order to gain additional insight into the effect s of fluorochemical exposure on workers’ health, an “episode of care” analysis was undertaken at the Decatur plant to screen for morbidity outcomes that may be associated with long-term, high ex posure to fluorochemicals (Olsen et al., 2001g; 2004c). An “episode of care” is a se ries of health care services prov ided from the start of a particu128OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

lar disease or condition until solution or resolution of that problem. Episodes of care were identified in employees’ health claims records using Clinical Care Groups (CCG) software. All inpatient and outpatient visits to health care providers, procedur es, ancillary services and prescription drugs used in the diagnosis, treatment, and management of over 400 diseases or conditions were tracked. Episodes of care were analyzed for 652 chemical employees and 659 film plant employees who worked at the Decatur plant for at least 1 y ear between January 1, 1993 and December 31, 1998. Based on work history records, employees were placed into different comparison groups: Group A consisted of all film and chemical plant worker s; Group B had employees who only worked in either the film or chemical plan t; Group C consisted of employees who worked in jobs with high POSF exposures; and Group D had employees who work ed in high exposures inthe chemical plant for 10 years or more prior to the onset of the stu dy. Film plant employees were considered to have little or no fluorochemical exposure, while chemi cal plant employees were assumed to have the highest exposures. Ratios of observed to expected episodes of care were calculated for each plant. Expected numbers were based on 3M’s employee population experience using indirect standard ization techniques. A ratio of the chemical plant’s obs erved to expected experience divi ded by the film plant’s observed to expected experience was calculated to provid e a relative risk ratio for each episode of care (RREpC). For each RREpC, 95% confidence interval s were calculated. Episode s of care that were of greatest interest were those wh ich had been reported in animal or epidemiologic literature on PFOS and PFOA: liver and bladde r cancer, endocrine disorders involving the thyroid gland and lipid metabolism, disorders of the liver and biliary tract, and reproductive disorders. The only increased risk of episodes for these conditi ons of a priori interest were for neoplasms of the male reproductive system and for the overall category of cancers and benign growths (which included cancer of the male repr oductive system). There was an in creased risk of episodes for the overall cancer category for all 4 comparison groups. The risk ratio was great est in the group of employees with the highest and l ongest exposures to fluorochemical s (RREpC = 1.6, 95% CI = 1.2 – 2.1). Increased risk of episodes in long-time, high-exposure employees also was reported for male reproductive cancers (RREpC = 9.7, 95% CI = 1.1 - 458). It should be noted that the confidence interval is very wide for male re productive cancers and the sub-categor y of prostate cancer. Five episodes of care were observed for reproductive cancer s in chemical plant employees (1.8 expected), of which 4 were prostate cancers (RREpC = 8.2, 95% CI = 0.8 - 399). One episode of prostate cancer was observed in film plant employees (3.4 exp ected). This finding should be noted because an excess in prostate cancer mortality was observed in the Cottage Grove plant mortality study when there were only 2 exposure categories (chemical division employees and non- chemical division employees). The update of the study sub-divided the ch emical plant employees and did not corroborate this finding when exposures were di vided into definitely exposed and probably exposed employees. There was an increased risk of episodes for neopla sms of the gastrointestinal tract in the high exposure group (RREpC = 1.8, 95% CI = 1.2- 3.0) and the long-term employment, high exposure group (RREpC = 2.9, 95% CI = 1.7 – 5.2). Most of the ep isodes were attributable to benign colonic polyps. Similar numbers of episodes were report ed in film and chemical plant employees. In the entire cohort, only 1 epis ode of care was reported for liver cancer (0.6 expected) and 1 for bladder cancer (1.5 expected ). Both occurred in f ilm plant employees. Only 2 cases of cirrhosis of the liver were observed (0.9 expected ), both in the chemical plant. Th ere was a greater risk of lower urinary tract infections in chemical plant employees , but they were mostly due to recurring episodes of care by the same employees. It is difficult to draw any conc lusions about these observations, given the small number of episodes reported.

129OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Chemical plant employees in the high exposure, long-term employment group were 2 1⁄2 times more likely to seek care for disorders of the biliary trac t than their counterparts in the film plant (RREpC = 2.6, 95% CI = 1.2 - 5.5). Eighteenepisodes of care were observe d in chemical plant employees and 14 in film plant workers. The sub-categories th at influenced this observation were episodes of cholelithiasis with acute cholecystitis and cholelith iasis with chronic or uns pecified cholecystitis. Most of the observed cases occurred in chemical plant employees. Risk ratios of episodes of care fo r endocrine disorders, which in cluded sub-categories of thyroid disease, diabetes, hyperlipidemia, and other endocrine or nutritional disorders, were not elevated in the comparison groups. Conditions which were not id entified a priori but which excluded the null hypothesis in the 95% confidence interval for the high exposure, long-term employment group included: disorders of the pancreas, cystit is, and lower urinary tract infections. The results of this study shoul d only be used for hypothesis ge neration. Although the episode of care design allowed for a direct comparison of work ers with similar demographics but different exposures, there are many limitations to this design. The limitations include: 1) episodes of care are reported, not disease incidence, 2) the data are difficult to interpret because a large RREpC may not necessarily indicate high ri sk of incidence of disease, 3) many of the risk ratios for episodes of care had very wide confidence intervals, thereby prov iding unstable results, 4) the analysis was limited to 6 years, 5) the utilization of health care services may reflect local medical practice patterns, 6) individuals may be counted more than once in th e database because they can be categorized under larger or smaller disease classifications, 7) episod es of care may include the same individual several times, 8) not all employees were included in the da tabase, such as those on long-term disability, 9) the analysis may be limited by the software use d, which may misclassify episodes of care, 10) the software may assign 2 different diag noses to the same episode, and 11) certain services, such as lab procedures may not have been reported in the database.

3.2.4 Other 3M Medical Surveillance Studies

3.2.4.1 Antwerp and Decatur Plants–Cross-S ectional and Longitudinal Studies A cross-sectional analysis of th e data from the 2000 medical surv eillance program at the Decatur 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 loca tions. Mean PFOA serum levels were 1.03 ppm for all male employees at the Antwerp plan t and 1.90 ppm for all male employees at the Decatur plant. Male production em ployees at the Decatur plant had significantly higher (p < .05) mean serum levels (2.34 ppm) thanthose at the Antwerp plant (1.28 ppm). Nonproduction employees at both plants had mean levels below 1 ppm. PFOA serum levels were higher than the PFOS serum values at both plants, especi ally 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 we re conducted to adjust for possi ble confounders that may affect the results of the clinical chemistry tests. Th e following variables were included: production job (yes or no), plant, age, BMI, cigarettes/day, drinks/day and years worked at the plant. A positive significant association was re ported between PFOA and choles terol (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 include d in the analyses, neither reached statistical significance in the cholesterol model, while PFOA remained significant (p = .02) in the triglycerides model. High-density lipoprotein (HDL) was ne gatively associated with PFOA (p = .04) and

130OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

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. BM I, 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 su rveillance results was performed to determine whether occupational exposure to fl uorochemicals 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, trig lycerides, alkaline phosphatase, gamma glutamyl transferase (GGT), aspartate ami notransferase (AST), alanine aminotransferase (ALT), total and direct bilirubin. Medical su rveillance data from 1995, 1997, and 2000 were analyzed using multivariable regression analysis. The pl ants were analyzed using 3 subcohorts that included those who participated in 2 or more me dical exams between 1995 and 2000. A total of 175 male employees voluntarily participated in the 2000 surveillance and at le ast 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 th e Antwerp plant increased between 1994/95 and 1997 and then decreased slightly be tween 1997 and 2000. At the Decatur plant, PFOA serum levels decreased between 1994/95 and 1997 and then increas ed between 1997 and 2000. When analyzed using mixed model multivariable regression and co mbining Antwerp and Decatur employees, there was a statistically significant positive associatio n between PFOA and 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 partic ular, 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 work ers. 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-sectiona l and longitudinal studi es 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 se veral 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 phosphata se, GGT, AST, and ALT va lues, 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 l ongitudinal study only a small number of employees participated in all 3 sampling periods (24%), diffe rent labs and analytical techniques for PFOA were used each year, and female employees could not be analyzed because of the small number of participants.

131OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

3.2.4.2 Cottage Grove Plant–Cross-sectional St udies of Clinical Chemistries and CCK 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, respectivel y) (Olsen et al., 1998b, 2000). The clinical chemistry parameters (chol esterol, hepatic enzyme s, 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, chol ecystokinin-33 (CCK) was also measured at the Cottage Grove plant. CCK levels were observed b ecause 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 year s. Cottage Grove has the highest serum PFOA levels of the 3 plants studied. Employees’ serum PFOA levels were stratifi ed into 3 categories (<1, 1- <10, and ≥ 10 ppm), chosen to provide a greater number of employees in the e 10 ppm category. As employees’ mean serum PFOA levels increased, no statisti cally significant (p < 0.05) abnormal liver function tests, hypolipidemia, or cholestasis were observed in any of the sampling years. Multivariable regression analyses controlling for potential confounders (age, al cohol consumption, BMI, and cigarettes smoked) yielded similar results. The authors also reported that renal function, bloo d glucose, and hematology measures were not associated with serum PFOA le vels; however, these datawere 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 e 10 ppm category. A statistically significant (p = .03) negative asso ciation 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 PF OA levels less than 10 ppm (0.6 and 5.6 ppm, respectively). A multiple regression model of th e 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 li mitation 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 ch anged substantially for ALT in 1997. Finally, different analytical methods were used to measure serum PFOA. Se rum PFOA was determined by electrospray highperformance liquid chromatography/mass spectro metry in 1997, but by thermospray in 1993 and 1995.
3M used data collected in their 2000 medical surveillance program to determine whether serum PFOA levels greater than 5 ppm were associated with changes in hepatic, lipid and thyroid function in workers (Olsen, et al., 2003f). Clinical chemis tries, including thyroid function tests, were performed for 131 male and 17 female workers at the plant. Serum samples were extracted and quantitatively analyzed for PFOA using HPLC/EMSS. All of the samples were above the lower limit of quantitation (LLOQ) (see Table 24 in Section 4.1.1 for biomonitoring data). Fifteen percent (n = 20) of male employees ha d serum concentrations that exceeded 5 ppm, and none of the female employees were above 5 ppm. Nu mber of years worked in the Chemical Division of the plant were not corre lated with PFOA serum measurements, but were correlated with specific production areas. When the male employees were separated into 3 groups based on serum PFOA levels (< 1 ppm, 1- 4.9 ppm, and >5 ppm), there were no stat istically significant (p < .05)

132OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

differences in mean lipid and hepa tic test results or inthyroid hormone levels, between the 3 groups both before and after adjusting for potential confounders (eg. BMI, smoking status, and alcohol consumption). The elimination of employees receivi ng cholesterol-re ducing drugs (n = 9) from the analysis did not alter these findings. Similar to fi ndings at the Decatur and Antwerp plants, triglyceride levels were higher in employees with th e highest PFOA serum con centrations, although not statistically significant (p < .05). In simple linear regression analys es, a weak negative association between T4 and serum PFOA concentrations was re ported (p = .07). Howeve r, all of the serum PFOA concentrations were within the T4 reference range, the statistical association explained minimal variation in the model (r2 = .03), there was no increase in serum TSH or T3 levels, and no negative association between free T4 levels and PFO A. No statistically signi ficant (p < 0.05) associations between PFOA and clinical chemistries or thyroid test resu lts for the small group of female employees was reported. An earlier medical surveillance study on worker s 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 re ported that animals exposed to PFOA develop hepatomegaly and alterations in lipid metabolis m, 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 stud ied to determine whether serum PFO A affected their cholesterol, lipoproteins, and hepatic enzyme s. Forty-eight workers who we re exposed to PFOA from 1985- 1989 were included in the study (96% participation rate). Sixty-fi ve employees who either volunteered or were asked to participate, were incl uded in the unexposed gr oup. These employees were assumed to have little or no PFOA exposure based on their job de scription. However, when serum levels were analyzed, it was noted that this group of workers had PFOA levels much greater than the general population. Therefore, in stead of job categories, total se rum 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), GGT, cholesterol, low-density lipoproteins (LDL), and HDL. All of the participants were placed into five categories of total seru m fluorine levels: <1 ppm, 1-3 ppm, >3 - 10 ppm, >10 - 15 ppm, and > 15 ppm. The range of the se rum fluorine values was 0 to 26 ppm (mean 3.3 ppm). Approximately half of the workers fell in to the > 1 - 3 ppm category, while 23 had serum levels < 1 ppm and 11 had levels > 10 ppm. There were no significant differenc es between exposure categories wh en 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 si gnificant differences among the exposure categories of total serum fluori ne for AST, ALT and GGT. However, increases in AST and ALT occurred with increasing total serum fluor ine levels in obese workers (BMI = 35 kg/m 2 ). This result was not observed when PFOA was measured directly in serum of workers in 1993, 1995, or 1997 surveillance data of employees of th e Cottage Grove plant (O lsen et al., 2000). Since PFOA was not measured directly and there is no exposure information provided on the employees (e.g. 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 re sults. 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 interes ting to compare the results of the

133OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

workers who were known to be exposed to PFOA to those who were origin ally 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).
Conclusions Several epidemiology and medical surveillance st udies have been conducted on workers employed at various APFO manufacturing sites in the U.S. Most of the studies were cross-sectional and focused primarily on males. A retros pective cohort mortality study de monstrated a statistically significant association between prostate cancer mortal ity and employment duration in the chemical facility of a plant that manufactures PFOA. However, in an update to this study in which more specific exposure measures were use d, a significant association for prostate cancer was not observed. Other mortality studies lacked adequate exposure da ta which could be linked to health outcomes. A study which examined hormone levels in workers reported an increase in estradiol levels in workers with the highest PFOA serum levels; however, th ese results may have been confounded by body mass index. Cholesterol and triglyce ride levels in workers were positively associated with PFOA exposures, which is inconsistent with the hypolipidem ic effects observed in rat studies. A statistically significant positive association was reported for PFOA and T3 levels in workers but not for any other thyroid hormones.

3.3 Mutagenicity Studies APFO was tested twice (Lawlo r, 1995; 1996) for its ability to induce mutation in the Salmonella – E. coli /mammalian-microsome reverse mutation assay. The tests were performed both with and without metabolic activation. A single positive res ponse seen at one dose level in S. typhimurium TA1537 when tested without metabolic activation was not reproducible. APFO did not induce mutation in either S. typhimurium or E. coli when te sted either with or w ithout metabolic activation. APFO did not induce chromosomal aberrations in human lymphocytes when tested with and without metabolic activation up to cytotoxic concentrations (M urli, 1996c; NOTOX, 2000). Sadhu (2002) reported that 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. Murli (1996b, 1996d) tested APFO tw ice for its ability to induce chromosomal aberrations in CHO cells. In the first assay, APFO induced both ch romosomal 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, whentested with metabolic activation, APFO induced significant in creases in chromosomal aberrati ons and in poly ploidy (Murli, 1996b).
APFO was tested in a cell transformation and cytotoxici ty assay conducted in C 3 H 10T 1⁄2 mouse embryo fibroblasts. The cell transformation was dete rmined as both colony transformation and foci transformation potential. There was no evidence of transformation at any of the dose levels tested in either the colony or foci assa y methods (Garry & Nelson, 1981).
APFO was tested twice in the in vivo mouse micronucleus assay. APFO did not induce any significant increases in micronuclei and was considered ne gative under the conditions of this assay (Murli, 1995, 1996a).
Conclusions APFO is not mutagenic. APFO di d not induce mutation in either S. typhimurium or E. coli when tested either with or without metabolic activation. APFO did not induce gene mutation when tested

134OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

with or without metabolic activat ion in the K-1 line of Chinese hamster ovary (CHO) cells in culture. APFO did not induce chromo somal aberrations in human lymphocytes when tested with and without metabolic activation up to cy totoxic 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 th e presence and absence of metabol ic 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 ne gative in a cell transf ormation assay in C 3 H 10T 1⁄2 mouse embryo fibroblasts and in the mouse micronucleus assay.

3.4 Acute Toxicity Studies in Animals Dean and Jessup (1978) reported an oral LD50 of 680 mg/kg and 430 mg/kg for male and female CD rats, respectively. Glaza (1997) reported an oral LD50 of greater than 500 mg/kg in male Sprague-Dawley rats and between 250 and 500 mg/kg infemales. Gabriel (1976c) reported an oral LD50 of less than 1000 mg/kg for male and female Sherman-Wistar rats. Rusch (1979) reported no mortality in male or female Sprague-Dawley rats following inhalation exposure to 18.6 mg/l for one hour. The dermal LD50 in New Zealand White rabb its was determined to be greater than 2000 mg/kg (Glaza, 1995). Kennedy (1985) determined a dermal LD50 of 4300 mg /kg for rabbits, 7000 mg/kg for male rats, and 7500 mg/kg for female rats. A PFO is an ocular irritant in rabbits when the compound is not washed from the eyes (Gabriel, 1976d), but is not an irritant in rabbits when washed from the eye (Gabriel, 1976a). Markoe (1983) found APFO to be a skinirritant in rabbits, while Gabriel (1976b) did not.
Conclusions 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 e xposure of 18.6 mg/l for one hour in rats. The dermal LD50 in rabbits was determined to be greater than 2000 mg /kg. The dermal LD50 is 4300 mg/kg for rabbits, 7000 mg/kg for male rats, and 7500 mg/kg for female ra ts. APFO is a primary ocular irritant in rabbits, while the data regarding poten tial skin irritancy are conflicting.

3.5 Repeat Dose Studies in Animals

3.5.1 Subchronic Studies in Non-Human Primates Goldenthal (1978b) administered rhesus monkeys (2/sex/group) doses of 0, 3, 10, 30 or 100 mg/kgday APFO by gavage for 90 days. Animals were observed twice daily and body weights were recorded weekly. Blood and ur ine samples were collect ed once during the contro l period, and at 1 and 3 months of the study for hematology, clinical ch emistry and urinalysis. Organs and tissues from animals that were sacrificed at the end of the study and from animal s that died during the treatment period were weighed, examined for gross pa thology and processed for histopathology. All monkeys in the 100 mg/kg-day group died during the study. The first death occurred during week 2; all animals were dead by week 5. Signs and symptoms which first appeared during week 1 included anorexia, frothy emesis which was sometimes brown in color, pale face and gums, swollen face and eyes, slight to severe decreased activ ity, prostration and body trembling. Three monkeys from the 30 mg/kg-day group died during the study; one male diedduring week 7 and the two females died during weeks 12 and 13. Beginning in week4, all four animals showed slight to moder135OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

ate and sometimes-severe decreased activity. One m onkey had emesis and ataxia, swollen face, eyes and vulva, as well as pallor of the face and gum s. Beginning in week 6, two monkeys had black stools and one monkey had slight to moderate dehydration and ptosis of the eyelids. No monkeys in the 3 or 10 mg/kg-day groups died during the study. One monkey in the 10 mg/kg-day group was anorexic during week 4, had a pale and swollen face in week 7 and had black stools for several days in week 12. Animals in the 3 mg/kg-day group occasio nally had soft stools or moderate to marked diarrhea; frothy emesis was also occasionally noted in this group. Changes in body weight were similar to the controls for animals from the 3 and 10 mg/kg-day groups. Monkeys from the 30 and 100 mg/kg-day groups lost body weight after week 1. At the end of the study, this loss was statistically significant for the one surviving male in the 30 mg/kg-day group (2.30 kg vs 3.78 kg for the control). The results of the urinalysis, and hematological and clinical chemistry analyses were comparable fo r the control and the 3 and 10 mg/kg-day groups at one and three months. At necropsy, the following changes in absolute and relative organ weight changes were noted: absolute and relative weight of the hearts in females from the 10 mg /kg-day group were significantly decreased; absolute brain weight of females from this same group we re also significantly decreased and relative group mean weight of the pituitary in males from th e 3 mg/kg-day group was significantly increased. The biological sign ificance of these weight changes is difficult to assess, as they were not accompanied by morphologic changes. In animals that died before the end of the study, one male and two females from the 30 mg/kg-day group and all animals from the 100 mg/kg-day group had marked diffuse lipid depletion in the adrenal glands. All males and females from the 30 and 100 mg/kg-day groups al so had slight to moderate hypocellularity of the bone marrow and moderate atrophy of lymphoid follicles in the spleen. One female from the 30 mg/kg-day group and all an imals in the 100 mg/kg-day group had moderate atrophy of the lymphoid follicles in the lymph nodes. Only one male in the 30 mg/kg-day group survived un til terminal sacrifice, and this male also had slight to moderate hypocellularity of the bone marrow and moderate atr ophy of lymphoid follicles in the spleen. No treatment related lesions were seen in the organs of animals from the 3 and 10 mg/kg-day groups.
The levels of PFOA in the serum and liver of th ese animals are presented below in Table 43. Individual values are presented so there are double entries for most dose levels.

Table 43 Levels of PFOA in the Serum and Liver of Surviving Rhesus Monkeys
Serum (μg/ml) Liver (μg/ml) Liver total (μg)
Dose Male Female Male Female Male Female

0 ND 1 0.05 0.07 3 5
3 53 65 3 7 250 350
3 48 50 ND ND ND ND
10 45 79 9 ND 600 ND
10 71 71 ND 10 ND 750
30 145 Dead 61 Dead 4000 Dead ND - Not Determined Source: Goldenthal 1978b
Under the conditions of this study the LOAEL is 3 mg/kg-day and no NOAEL was established.

136OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Thomford (2001a, b) conducted a range-finding and a 6-month toxicity study in male cynomolgus monkeys. In the range-finding study, Thomford (2001a) administered male cynomolgus monkeys an oral capsule containing 0, 2, or 20 mg/kg-day APFO for 4 weeks. There were 3 monkeys in the 2 and 20 mg/kg-day groups and one monkey in the c ontrol group. The monkey s weighed 2.1 to 3.6 kg at the start of treatment. Animals were observed twice daily for mortality and moribundity and were examined at least once daily for signs of poor hea lth or abnormal behavior. Body weights were recorded weekly and food consumption was assessed qualitatively. The monkeys were fasted overnight and blood samples were collected one week prior to the start of the study and on day 30 for clinical hematology and clinical chemistry analyses, and hormone and PFOA level. Blood for clinical chemistry was also collected from each animal on day 2 (approximately 24 hours after the first dose). Samples were analyzed for estradiol, estron e, estriol, thyroid stim ulating hormone, total and free triiodothyronine, and to tal and free thyroxin. At scheduled necropsy, samples of the right lateral lobe of the liver were collected from each animal and analyzed for palmitoyl CoA oxidase activit y. Representative samples of liver, right and left testes, and pancreas were collected from each anim al for cell proliferation evaluation using proliferation cell nuclear antigen. Bile was collected fr om each animal for bile acid determination. A sample of liver was collected from each animal for PFOA concentration analysis. The adrenals, liver, pancreas, spleen, and testes from each anim al were examined microscopically, and the remaining tissues were preserved for possible future examination.

All animals survived to scheduled s acrifice. There were no clinical signs of toxicity in the treated groups and there was no effect on body weight. Low or no food cons umption was observed for one animal given 20 mg/kg-day. There were no effects on estradiol, estriol, thyroid stimulating hormone, total and free triiodothyronine , and total and free thyroxin. Estr one levels were notably lower for males given 2 and 20 mg/kg-da y APFO. There was no evidence of peroxisome proliferation or cell proliferation in the liver, test es or pancreas of tr eated monkeys. No adverse effects were noted in either gross or clinical pathology studies. In the 26-week study, male cynomolgus monkeys were administered APFO by oral capsule at doses of 0, 3, 10 or 30 mg/kg-day for 26 weeks (Thomfo rd 2001b; Butenhoff et al., 2002). At study initiation the monkeys weighed 3.2 to 4.5 kg. There we re 4 monkeys in the 3 mg/kg-day group and 6 monkeys in each of the other groups. Dosing of animals in the 30 mg/kg-day dose group was stopped from days 11–21 because oftoxicity. When dosing was resumed on day 22, animals received 20 mg/kg-day and this group was designat ed the 30/20 mg/kg-day gr oup. At the end of the 26-week treatment period, 2 animals in the cont rol and 10 mg/kg-day groups were observed for a 13-week recovery period. Animals were observed twice daily for mortality and moribundity and were examined at least once daily for signs of poor health or abnormal behavior. Ophthalmic examinations were done before initiation of treatment and during weeks 26 and 40. Body weights were recorded weekly and food consumption was assessed qualitatively. Blood and ur ine samples were collect ed for clinical hematology, clinical chemistry, and urin alysis before the start of tr eatment and on days 11, 31, 63, 91, 182, 217, 245 and 275. Blood samples were also taken for hormone determinations; samples were analyzed for estradiol, estrone, estriol, thyroi d stimulating hormone, total and free triiodothyronine, total and free thyroxin, testoste rone, and cholecystokinin (CCK). Blood, urine and feces were collected during week 2 and every 2 weeks thereafte r during treatment and recovery for PFOA concentration analyses. At scheduled necropsy, liver sample s were taken for determination of PFOA levels. The right lateral lobe of the liver was collected from each animal for palmitoyl CoA oxidase activity analyses, and representative samples of liver, right and left test es, and pancreas were collected from each animal for cell proliferation evaluation us ing proliferation cell nuclear antigen. All available bile was col137OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

lected for bile acid determination. Weights of th e adrenal glands, brain, ep ididymis, kidney, liver, pancreas, testis, and thyroid with parathyroid we re recorded. The following tissues were collected for histopathology: adrenal (2), aorta, brain, cecum, colon, duodenum, epididymis (2), esophagus, eyes [preserved in Davidson’s fixative (2)], femur with bone marrow (articular surface of the distal end), gallbladder, heart, ileum, jejunum, kidneys (2), lesions, liver, lung, mesenteric lymph node, mammary gland, pancreas, pituitary, prostate, rectum, salivary gland [mandibular (2)], sciatic nerve, seminal vesicle (2), skeletal muscle (thigh), ski n, spinal cord (cervical, thoracic, and lumbar), spleen, sternum with bone marrow, stomach, testis [(2) preserved in Bouin’s solution], thymus, thyroid (2) with parathyroid, trachea and urinary bladder. Two animals, one male from the 30/20 mg/kg-day dose group and one male from the 3 mg/kg-day dose group, were sacrificed in moribund condition during the study. The male in the 30/20 mg/kgday dose group was sacrificed on day 29. This animal exhibited signs of hypoactivity, weight loss, few or no feces, low or no food consumption and the entire body was cold to the touch before death. Necropsy revealed esophageal and ga stric lesions that were indicative of an injury that occurred during dosing and liver lesions that were presumed to be treatmentrelated. The animal from the 3 mg/kg-day dose group was sacrificed on day 137. This animal showed clinical signs of limited use and paralysis of the hind limbs, ataxia and hypo active behavior, few feces and no food consumption. The cause of death was not determined, but APFO treatment could not be ruled out. Males given 30 mg/kg-day from days 1-11 had clin ical signs of few feces and low food consumption and they lost weight during week 1 of trea tment. Based on these signs, treatment was stopped on day 11 and was not resumed until day 22. When treatment was resumed, the dose was lowered to 20 mg/kg-day; this group was then designated the 30/20 mg/kg-day group. Of the remaining animals in this group, only 2 tolerate d this dose level for the remaini ng 23 weeks of treatment. Treatment of three males given 30/20 mg/kg-day was halted on days 43 (week 7), 66 (week 10), and 81 (week 12) respectively. Clinical si gns in these animals included thin appearance, few or no feces, low or no food consumption, and weight loss. Th e animals appeared to recover from compoundrelated effects within 3 weeks after cessation of treatment. Mean body weight changes were notably lower during weeks 1 and 2 for males receiving 30 mg/kgday. During week 2, this change wa s statistically significant. Treatment was stopped on day 11 and when it was resumed at 20 mg/kg-day on day 21, mean body weight change s were significantly lower than controls during weeks 7, 9 and 24. Ov erall mean body weight changes through week 27 were notably lower for the males in the 30/20 mg/kg-day group. There was an increased incidence of low or no food consumption for animals in the 30/20 mg/kg-day group that was considered to be treatment related. There were no consistent or clearly dose-related eff ects on estrone, estradiol, estriol, testosterone, or CCK levels in treated groups over time. Similarl y, thyroid stimulating hormone and free and total thyroxine levels remained relatively constant and were not si gnificantly altered throughout the study. Mean individual values for free and total trii odothyronine levels were statistically significantly decreased during the dosing period at the 30/20 mg/kg/day dose group compared to controls, although the biological si gnificance of this decrease is unclear. At terminal sacrifice at 26 weeks, there were statistically signifi cant increases in mean absolute liver weights and mean liver-to-body weight percen tages in all dose groups. In addition, there was a positive dose-response trend towards an increased relative liver-to-body weight , with statistical significance at the high dose group of 30/20 mg/kg/day. The increased liver weights were considered to be treatment-related. The increase d liver weight was thought to be due, in part, to hepatocellular hypertrophy (as demonstrated by decreased hepatic DNA content) which in turn may be due to mitochondrial proliferation (as demonstrated by in creased succinate dehydrogenase activity).

138OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Since administration of APFO to rats results in li ver, Leydig cell and panc reatic acinar cell tumors, Butenhoff et al (2002) specifically looked for mark ers of tumor formation in the monkeys. In the liver, there was only a two-fold increase in hepatic palmitoyl CoA oxidase activity in the 30/20 mg/kg-day group, which is consistent with reports fo r other species that are not particularly responsive to PPAR • -agonists. Replicative DNA synthe sis in the liver, an indi cation of cell proliferation, was not altered in the treated animals. Similarly, it has been proposed that changes associated with the pancreatic acinar cell tumors in rats includ e increased serum CCK concentrations and indications of cholestasis, including al kaline phosphatase, bilirubin, and bi le acids; none of these changes were noted in the cynomolgus monkeys. Finally, in the rat, it has been proposed that the Leydig cell tumors are due to a sustained increase in estradio l resulting from aromatase induction. In the treated cynomolgus monkeys there were no significant changes in estradiol, estriol, ortestosterone. In addition, there was no change in replicative DNA synthesis in the pancreas or testes. At the recovery sacrifice, there were no treatme nt-related effects on term inal body weights or on absolute or relative organ weights indicating that the liver weight ch anges seen at terminal sacrifice were reversible over time. There were no treatmen t-related macroscopic or microscopic changes at the recovery sacrifice.
Serum and liver concentrations of PFOA did not in crease in a dose-related manner. This may have been due to saturation or attaining steady state levels in the first several weeks of the study. In addition, there was a great deal of vari ability in the PFOA levels. This may have been due to the method of dosing (by capsule), the timing of dosing rela tive to blood sample collection and gall bladder emptying, or the analytical method ( + 30% for interassay, intra-assay and system). Since steady state appeared to have been re ached by 4-6 weeks of dosing, the st udy authors calculated the mean serum levels for the period following 6 weeks; steady state serum levels were 77 + 39, 86 + 33 and 158 + 100 μg/ml for the 3, 10 and 30/20 mg/kg-day groups , respectively. In the control animals, 2/3 of the serum samples contained PFOA and averaged a level of 0.203 + 0.154 μg/ml. There was not a statistically significant difference in the mean serum levels of the 3 and 10 mg/kg-day groups, but the mean serum level in the 30/20 mg/kg-day group was significantly hi gher than the 3 and 10 mg/kg-day groups. At terminal sacrifice, the levels of PFOA in the liver were similar in the 3 and 10 mg/kg-day groups, and ranged from 6.29 - 21.9 μg/ g. The two monkeys in the 30/20 mg/kg-day group had liver concentrations of 16 and 83.3 μg/g. After the recove ry period, the serum and liver PFOA levels in the 10 mg/kg-day group had returned to baseline. Under the conditions of the study, the LOAEL was 3 mg/kg-day (incre ased liver weight and possibly mortality) and a NOAEL was not established.

Conclusions Repeat-dose studies have been conducted in non- human primates. In a 13-week study with Rhesus monkeys, exposure to doses of 30 mg/kg-day or higher resulted in death. C linical signs of toxicity were noted at doses as low as 3 mg/kg-day. Under the conditions of this study the LOAEL is 3 mg/kg-day and no NOAEL was established. In a 6-month study of male cynomolgus monkeys, there was a steep dose response curve for mortality. In creases in liver weight were noted at doses as low as 3 mg/kg-day, but there was no evidence of pe roxisome proliferator-activated receptor alpha activity (PPAR α ). Under the conditions of the study, the LOAEL was 3 mg/kg-day (increased liver weight and possibly mortality) and a NOAEL was not established.

139OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

3.5.2 Subchronic Studies in Rodents Christopher and Marias (1977) administered ChR-CD mice (5/sex/group) diet ary concentrations of0, 30, 100, 300, 1000, 3000, 10,000, or 30,000 ppm of APFO for 28 days. The animals were observed daily and body weights an d food consumption were recordedweekly. At necropsy, the organs were weighed, examined for gross patholog y and preserved for histopathology. All animals in the 1000 ppm and higher groups died before the e nd of day 9. All animals in the 300 ppm group died within 26 days except one male. One animal in each of the 30 and 100 ppm groups died prematurely. Clinical signs were observed in mice exposed to 100 ppm and higher doses of APFO. At 100 ppm some animals exhibited cyanosis on days 10 and 11 of testing, but appeared normal throughout the rest of the study. Animals fed 300 ppm exhibite d roughed fur and muscular weakness as well as signs of cyanosis after 9 days of treatment. Anim als fed 1000 ppm exhibited similar effects after 6 days and those receiving 3000 ppm or greater doses exhibited eff ects after 4 days. There was a dose-related reduction in mean bo dy weight in all treated groups . Relative and absolute liver weights were increased in mice fed 30 ppm or more APFO. Treat ment-related changes were observed in the livers among all APFO treated animals incl uding enlargement and/or discoloration of 1 or more liver lobes. Histopathologic examinati on of all surviving treated mice revealed diffuse cytoplasmic enlargement of hepatocytes throughout the liver (panlobular hype rtrophy) accompanied by focal to multifocal cytoplasmic lipid vacuoles of variable size which were random in distribution. Metrick and Marias (1977) administered ChR-CD ra ts (5/sex/group) dietary concentrations of 0, 30, 100, 300, 1000, 3000, 10,000, or 30,000 ppm APFO for 28 days. The animals were observed daily and body weights and food consump tion were recorded weekly. Atnecropsy, the organs were weighed, examined for gross pathology and preser ved for histopathology. All animals in the 10,000 and 30,000 ppm groups died before th e end of the first week. Gro ss pathologic examination revealed white foci in the cortex and medulla inthe kidneys of a 10,000 ppm female. Pelvic dilation was evident in the kidneys of a control male and a 30,000 ppm female. There were no premature deaths or unusual behavioral r eactions in the other groups. Food consumption was reduced in the 1000 and 3000 ppm groups. Body weight gain was reduced as dose increas ed. The reduction in body weight gain was statistically significant fo r males at 1000 ppm and ma les and females at 3000 ppm. Absolute liver weights were increased in males fed 30 ppm or more and females fed 300 ppm or more. Treatment-related morphologic changes were observed in the livers of all test animals. Focal to multifocal cytoplasmic enlargement of hepatocytes in the centrilobular to midzonal areas was noted in animals fed 30 to 300 ppm, and multifocal to diffuse enlargement of hepatocytes throughout the liver lobules (panlobular) was noted in animals fed 1000 ppm or higher. The hypertrophy of hepatocytes was accompanied by aci dophilic degeneration and/or necros is of scattered liver cells with no lobular distribution. The severity and de gree of tissue involvement were more pronounced in males than in females. Goldenthal (1978a) administered ChR-CD rats (5/sex/group) dietary levels of 0, 10, 30, 100, 300, and 1000 ppm APFO for 90 days. These dose le vels are equivalent to 0.056, 1.72, 5.64, 17.9, and 63.5 mg/kg-day in males, and 0.74, 2.3, 7.7, 22.36 and 76.47 mg /kg-day in females. Animals were observed twice daily and body wei ght and food consumption were recorded weekly. Blood and urine samples were collected duri ng the pretest period and at 1 and 3 months of the study for hematology and clinical chemistry and urinalysis. At necropsy, the orga ns from the control, 100, 300, and 1000 ppm groups were weighed and examined for hist opathologic lesions; livers from the 10 and 30 ppm groups were also examined microscopically.

There were no treatment-related ch anges in behavior or appearance. One female in the 100 and one female in the 300 ppm group died dur ing collection of blood. These d eaths were not considered to be treatment related. All other animals survived until scheduled sacrifice. There was a decrease in body weight gain for male rats at the 300 and 1000 ppm dose level. At 13 weeks, mean body weight

140OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

of males in the 1000 ppm group was significantly less th an that of controls. There were no treatment related effects on the hematologic, biochemical or urine parameters. Relative kidney weights were significantly increased in males in the 100, 300, and 1000 ppm groups. However, absolute kidney weights were comparable among groups, and there were no histopathological lesions. Absolute liver weights were significantly increased in males in the 30, 300 and 1000 ppm groups and in females in the 1000 ppm group. Relative liv er weights were significantly increased in males in the 300 and 1000 ppm groups and in females in the 1000 ppm group. Discoloration on the surface of the liver was observed in male rats in the 1000 ppm group. Hepatocellular hypertrophy (focal to multifocal in the centrilobular to midzonal regions) was observed in 4/5, 5/5, and 5/5 male s in the 100, 300, and 1000 ppm gr oups, respectively. Hepatocyte necrosis was observed in 2/5, 2/5, 1/5, and 2/ 5 males in the 30, 100, 300, and 1000 ppm groups, respectively. Under the conditions of this study, the LOAEL for males is 30 ppm (1.72 mg/kg-day) based on liver effects and the NOAEL is 10 ppm (0.56 mg/kg-day); the LOAEL for females is 1000 ppm (76.5 mg/kg-day) and the NOAEL is 300 ppm (22.4 mg/kg-day). In a dietary study, male ChR-CD ra ts (45-55 per group) were administ ered concentrations of 1, 10, 30, or 100 ppm APFO for 13 weeks (Palazzolo 1993; Perkin s et al., 2004). These doses are equivalent to 0.06, 0.64, 1.94, and 6.50 mg/kg-day. Two contro l groups (a nonpair-fed control group and a control group pair-fed to the 100 ppm dose group) were fed basal diet during that period. Following the 13-week exposure period, 10 animals per group were fed basal diet for an 8-we ek recovery period. The animals were observed twice daily for c linical signs of toxicity, and body weights and food consumption were recorded weekly. Food cons umption was recorded daily for the pair-fed animals. A total of 15 animals per group were sacrif iced following 4, 7, or 13 weeks of treatment; 10 animals per group were sacrificed after 13 weeks of treatment and following the 8 week recovery period. Serum samples collected from 10 animals per group at each sche duled sacrifice during treatment and from 5 animals per group during recovery were analyzed for estr adiol, total testosterone, luteinizing hormone, and PFOA. The level of palmitoyl CoA oxidase was analyzed from a section of liver that was obtained from 5 animals per group at each scheduled sacrifice. Weights of the brain, liver, lungs, testis, seminal vesicle, prostate, coagulating gland, and urethra were recorded, and these tissues were examined histologically. In addition, the brain, liver, lungs, testis, seminal vesicle, and prostate were pr eserved in glutaraldehyde for el ectron microscopic examination. In the analysis of the data, animals in groups exposed to 1, 10, 30, and 100 ppm APFO were compared to the control animals in the nonpair-fed group, while the data from the pair-fed control animals were compared to animals exposed to 100 ppm APFO. No treatment-related clinical signs were noted. At 100 ppm, significant reductions in body weights were seen compared to the pair-fed control group during week 1 and the nonpair-fed control group during weeks 1-13. Body weight data in the other dosed-groups were comparable to controls. At 100 ppm, mean body weight gains were significantly higher than the pair-fed control group during week 1 and significantly lower than the nonpair-fed control group during weeks 1-13. At 10 and 30 ppm, mean body weight gains were significantly lower than the nonpair-fed control gr oup at week 2. These differences in body weight and body weight gains were not observed during the recovery period. Animals fed 100 ppm consumed significantly less food during weeks 1 and 2, when compared to the nonpair-fed control group. Overall, there was no significant difference in food consumption. There were no significant differences among the groups for any of the horm ones evaluated in the serum although there appeared to be some indication of elevatedestradiol for the 100 ppm group at week 5. Significant increases in absolute and relative liver weights and hepatocellular hypertrophy were observed at weeks 4, 7, or 13 in the 10, 30 and 100 ppm groups. There was no evidence of any degenerative changes or abnormalities associated with the hypertrophy. Hepatic palmitoyl CoA oxidase

141OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

activity was significantly increased at weeks 4, 7, and 13 in the 30 and 100 ppm groups. At 10 ppm, hepatic palmitoyl CoA oxidase activity was signi ficantly increased at week 4 only. During recovery, however, none of the liver effects were observ ed, indicating that these treatment-related liver effects were reversible.
Serum levels from the male rats in this study are shown below in Table 44.

Table 44 Summary of Serum Values (μg/ml ) in Male Rats Treated with APFO for 90 Days
Dose Group (mg/kg-day)

Week 0.06 0.64 1.94 6.50

4 6.5 ± 1.0 a (8) 55 ± 8.1 (9) 104 ± 14 (8) 159 ± 30 (10)
7 7.5 ± 1.3 (9) 46 ± 16 (10) 87 ± 28 (10) 149 ± 35 (10)
13 7.1 ± 1.2 (10) 41 ± 13 (10) 70 ± 16 (10) 138± 34 (10)
21 1.2 (1) 1.1± 1.3 (3) 1.6 ± 0.9 (3) 2.5± 0.9 (2)

a Mean +SD, (N) Source Perkins et al., 2004
Under the conditions of this study, the LOAEL is 10 ppm (0.64 mg/kg-day) based on increases in absolute and relative liver weights with hepatocellular hypertrophy. Th e NOAEL is 1.0 ppm (0.06 mg/kg-day).
Conclusions Repeat-dose studies in rats and mice demonstrated th at the liver is the primary target organ. Due to gender differences in elimination, adult male rats exhibit effects at lower administered doses than adult female rats. In a 90-day dietary study of APFO in rats , the LOAEL was 10 ppm (0.64 mg/kgday) based on increases in absolu te and relative liver weights w ith hepatocellular hypertrophy, and the NOAEL was 1.0 ppm (0.06 mg/kg-day).

3.6 Carcinogenicity Studies in Animals

3.6.1 Chronic and Carcinogenicity Studies in Rats The chronic toxicity and carcinogenic potential of PFOA has been investigatedin two dietary studies in rats. In the first study (Sibinski, 1987), groups of 50 male and 50 female Sprague-Dawley (Crl:CD BR) rats were fed diets containing 0, 30 or 300 ppm APFO for two years. Groups of 15 additional rats per sex were fed 0 or 300 ppm APFO and evaluated at the one year interim sacrifice. In males. the mean test article consumpti on was 1.3 and 14.2 mg/kg-day for the 30 and 300 ppm groups, respectively; in females, the mean testarticle consumption was 1.6 and 16.1 mg/kg-day for the 30 and 300 ppm groups, respectiv ely. All animals were observed daily throughout the two year

142OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

dosing period. Body weights and feed consumption were recorded once per week for the first six months, and then once every two weeks for the rema inder of the study. Clinical pathological examinations including hematology, serum chemistry and urinalysis were conducted on samples obtained from 15 rats per sex from each group at 3, 6, 12, 18 and 24 months. Macroscopic postmortem examinations were performed on all animals thatdied during the study and those which were terminated at the one year interim and two year necr opsies. The weights of th e kidneys, liver, testes, brain, heart, spleen, adrenal glands and uterus were recorded for 15 randomly selected rats/sex at the interim termination from both the control an d high-dose groups, and from the control and both treated groups at the two year necropsy. Microscopic evaluation wa s performed on all tissues from all of the control and high-dose rats. There was a dose-related decrease in body weight gains in the male rats and to a lesser extent, in the female rats as compared to the controls; the decr eases were statistically significant in the high-dose groups of both sexes. In the high-dose (300 ppm) males, there was a 21% decrease in body weight gain compared to the control males by week 6 a nd the body weight gains were decreased over 10% compared to the control males through 66 weeks of the study. This difference was statistically significant until week 98. In the lo w-dose (30 ppm) males, a 5% decrease in body weights was observed at week 6, however, there was little additional decrease th ereafter. Mean body weights were only slightly decrease in the treated females comp ared to the control values through the first 18 months of the study but reached a maximum of 11% decrease in the high-dose females at 92 weeks. The body weight changes are treatment related si nce feed consumption was actually increased (rather than decreased). There were no differences in mortality between th e treated and untreated groups; the survival rates at the end of 104 weeks for the male control, lo w-, and high-dose groups were 70%, 72% and 88%, respectively; in females, the surv ival rates were 50%, 48% and 58% for the control, 30 and 300 ppm groups, respectively. The onl y clinical sign observed was a dose-re lated increase inataxia in the female rats; which was most co mmonly associated with moribund animals; the incidences in the control, low- and high-dose groups were 4%, 18% and 30%. Significant decr eases in red blood cell counts, hemoglobin concentrations and hematocrit values were observed in the high-dose male and female rats as compared to control values. Clinic al chemistry changes incl uded slight (less than 2- fold), but significant increases (P < 0.05) in alanine aminotransferase (ALT), aspartate aminotransferase (AST) and alkaline phosphatase (AP) in bo th treated male groups fr om 3-18 months, but only in the high-dose males at 24 months. Slight (up to about 10%) increases in absolute or relative liver and kidney weights were not ed in both high-dose male and female rats at the 1 year interim sacrifice and at the terminal necropsy; however, only the relative liver weight (vs. body weight or brain weight) increases in the high-dose males we re statistically si gnificant (P<0.05). Histologic evaluations showed lesions in the liver, testis and ovary. In the liver, the increased incidence of lesions reached statistical significance only in the hi gh-dose male group. At the 1 year interim sacrifice, diffuse hepatomegalocytosis (1 2/15 animals), portal mononuc lear cell infiltration (13/15 animals) and hepatocellular necrosis (6/15 animals) were seen in the high-dose males while incidences in the control group were, 0/15, 7/15 and 0/15, respectively. Hepa tocellular vacuolation was seen in 11/15 high-dose females as compared to an incidence of 5/15 in the control group. At the 2-year sacrifice, megalocyto sis was found at an incidence of 0%, 12% and 80% in the males, and 0%, 2% and 16% in the females from the c ontrol, low-, and high-dose groups, respectively. Hepatic cystoid degeneration, a condition characteri zed by areas of multilocular microcysts in the liver parenchyma, was observed in 14% and 56% of the low- and high-dose males, as compared to a control incidence of 8%. The inci dence of hyperplastic nodules, a lo calized proliferation of hepatic parenchymal cells, was slightly in creased in the high-dosed males w ith an incidence of 6% as compared to 0% in the control males.

143OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

At the one-year sacrifice, tes ticular masses were found in 6/50 high-dose and 1/50 low-dose rats, but not in any of the co ntrols. Furthermore, marked asperm atogenesis was found in 2/15 high-dose males but none in the control males. At the 2-year sacrifice, vascular mine ralization of the testes occurred in 18% of the high-dosed male and 6% of the low-dosed males, but was not seen in the controls. These testicular effects reached statistical significance in the high-dose group. A statistically significant, dose-related increase in the incidence of ovarian tubular hyperplasia was found in female rats at the 2-year sacrifice. The incidence of this lesion in the control, low-, and high-dose groups was 0%, 14%, and 32%, respectively. Th e biological significance of this effect at the time of the initial evaluation was unknown, as th ere was no evidence of progression to tumors. Recently, however, slides of the ovaries from that study were re-evaluated, with particular emphasis placed on the proliferative lesions of the ovary (Mann and Frame, 2004). Using more recently published nomenclature, the ovarian lesions were dia gnosed and graded as gonadal stromal hyperplasia and/or adenomas, which corresponded to the dia gnoses of tubular hyperplasia or tubular adenoma by the original study pathologist. Th e data are summarized in Table 45. No statistically significant increases in hyperplasia (total num ber), adenomas, or hyperplasia/ad enoma combined were seen in treated groups compared to controls. There was so me evidence of an increas e in size of stromal lesions observed at the 300 ppm group; however, adenomas occurred ingreater incidences in the control group than in either of the treated groups. Resu lts of this follow-up evaluation indicated that rats sacrificed at the one-year interi m sacrifice, as well as rats that died prior to the interim sacrifice were not considered at risk for tumor development.

Table 45 Incidence of Ovarian Stromal Hyperplasia and Adenoma in Rats
Group Control 30 ppm 300 ppm
No. examined 45 47 46 Hyperplasia (Total) 8 16 15
Grade 1 6 7 5
Grade 2 2 3 1
Grade 3 0 5 6
Grade 4 0 1 3
Adenoma 4 0 2Adenoma and/or Hyperplasia 12 16 17 From Mann and Frame (2004) The non-cancer effects in this study, the LOAEL for male rats is 300 ppm and the NOAEL is 30 ppm based on a decrease in body weight gain, increas e in liver and kidney weights and toxicity in the hematological and hepatic systems. The LOAEL for female rats is 300 ppm based on a decrease in body weight gain and hematologic effects and the NOAEL is 30 ppm. At the termination of the study, th ere was a significant increase (P<0.05) in the incide nce of testicular (Leydig) cell adenomas in the high-dose male rats. The incidence of the Leydig cell tumors (LCT) in the control, low- and high-dose groups was 0/50 (0%), 2/ 50 (4%) and 7/50 (14%), respectively. The increase was also statistically significant when compared to the historical control incidence of 0.82% observed in 1,340 Sprague-Dawley c ontrol male rats used in 17 carcinogenicity studies (Chandra et al., 1992). The spontaneous in cidence of LCT in 2-year old Sprague-Dawley rats in other studies was repor ted to be approximately 5% (c ited in Clegg et al., 1997).

144OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

The report also indicated a signif icant increase (P<0.05) in the in cidence of mammary fibroadenomas in both groups of female rats. The incidenc e of the mammary fibroa denoma was 21% (10/47), 40% (19/47) and 43% (21/49) in the control, 30, and 300 ppm gr oups, respectively. The increase was also statistically si gnificant when compared to the hist orical control in cidence of 19.0% observed in 1,329 Sprague-Daw ley control female rats used in 17 carcinogenicity studies (Chandra et al., 1992). The investigators did not consider the mammary fibroade nomas to be treatment related on the basis of the historical control incidence (24%) from a study of 181 female rats terminally sacrificed at 18 months, which is considered an inappropriate hist orical reference. When the mammary fibroadenoma incidences were compared to the historical control incidence (37%) in 947 female rats in the Haskell Laboratory, however, there does not appear to be any compound related effect (Sykes,1987). The mammary gland findings were re-examined recently by a Pathology Working Group (Hardisty, 2005) using diagnostic criteria an d nomenclature of the Society of Toxicological Pathologists (Mann et al., 1996). The PathologyWorking Group (PWG) concluded th at there were no statistically significant differences in the incidence offibroadenoma, adenocarcinoma, total benign neoplasms or total malignant neoplasms of the mamm ary glands between control and treated animals using Fischer’s Exact Test for pair-wise comparis ion. There was also no significant difference in combined benign and malignant neoplasms between control and treated groups. The primary difference between the original reported findings and the PWG results involved findings initially reported as lobular hyperplasia which the PWG classified as fibroadenoma, mostly in the control group. According to the PWG, the incidence of mammary fi broadenoma in the control, low- and high-dose groups were: 32% (16/50), 32% (16/50), and 40% (20/50), respectively. The induction of Leydig cell tumors was confirmed in a follow-up 2-year mechanistic study of PFOA toxicity in male Sprague-Dawley rats at a dietary level of 300 ppm (C ook et al., 1994; Biegel et al., 2001). There were 156 animals in the trea tment group and 80 animals in the control group. Cage-side examinations were conducted at le ast once daily throughout the study. Rats were weighed once a week during the first 3 months and once every other week for the remainder of the study. Rats were euthanized at interim time poi nts of 1, 3, 6, 9, 12, 15, 18, and 21 months. At each time point, the liver and testis from 6 rats/group were weighed and evaluate d for cell proliferation. Another 6 rats/group were selected for peroxiso me proliferation, and 10 rats/group for serum hormone (estradiol, testosterone, LH, FSH, and prolactin) analysis. All rats surviving the 24-month test period were necropsied for microscopic examinati on of various organs: e .g., kidneys, liver, testes, brain, heart, spleen. In the treated group, relative liver weights and hepatic β -oxidation activity were statistically significantly increased at all of the sampling time points wh en compared to the controls. Absolute testis weights were increased only at 24 months. No hepa tic or Leydig cell prolif eration was observed at any of sampling times. There were no significant differences in serum te stosterone, FSH, LH, or prolactin in the PFOA-treated rats when compared to the controls. There were, however, significant increases in serum estradiol concentrations in the treated rats at 1, 3, 6, 9, and 12 months. There was a significant increase in the incidence of LCT in the treated rats 11% (8/76) as compared to the controls (0%, 0/80). In addition, the treated group had a signi ficant increase in the incidences of liver adenomas and pancreatic acinar cell tumo rs (PACT). The incidences of liver adenomas in the control and treated groups were 3% (2/80) and 13% (10/76), re spectively, whereas those for the pancreatic acinar cell adenomas were 0% (0/80) and 9% (7/76). There was one pancreatic acinar cell carcinoma in 76 of th e treated rats and none in 80 controls . The incidence of combined pancreatic acinar cell adenoma/carcinoma in the treated rats (11%, 8/76) was significantly increased as compared with the controls (0%, 0/80).

145OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

In the first carcinogenicity study (S ibinski, 1987), there was no reporte d increase in the incidence of PACT, and the incidence of pancreatic acinar hyper plasia in the male rats was 0/33, 2/34, and 1/43 in the control, 30 and 300 ppm groups, respectivel y. To resolve this discre pancy, the histological slides from both studies were reviewed by independent pathologists. This review of the microscopic lesions of the pancreas in the two studies indicates that PFOA pr oduced increased incidences of proliferative acinar cell lesions of the pancreas in the rats of both st udies at the dietary concentration of 300 ppm. The differences observed were quantitative rather than qualitiative; more and larger focal proliferative acinar cell lesions and greater tendency for progression of lesions to adenoma of the pancreas were observed in the second study compared to the fi rst study. The difference between pancreatic acinar hyperplasia (reported in Sibi nski, 1987) and adenoma (reported in Cook et al., 1994; Biegel et al. 2001) in the rat is a reflection of arbitrary diagnostic criteria and nomenclature by different pathologists. The basis for the quantit ative difference in the lesions observed in not known but was believed to be due most likely to diffe rence in the diets used in the two laboratories (Frame and McConnell, 2003).
Conclusions Two carcinogenicity studies of PFOA have show n that PFOA induced liver adenomas, Leydig cell adenomas, and pancreatic acinar cell tumors in male Sprague-Dawley ra ts. PFOA has also been shown to promote liver carcinogene sis in rodents (Abdellatif et al.,1991; Nilsson et al., 1991). The evidence for mammary fibroadenomas in the female rats is equivocal since the incidences were comparable to some historical background incidences and a recent Pathology Working Group (Hardisty, 2005) concluded that ther e were no statistically significant differences in the incidence of fibroadenoma or other neoplasms of the mammary glands between control and treated animals.

3.6.2 Mode of Action Studies The modes of carcinogenic action of PFOA-induced LCTs and PACT have not been fully elucidated (Klaunig et al., 2004). However, as describe d below, there is ample evidence that the liver tumors in rats are due to a PPAR α -agonist mode of action (MOA).

3.6.2.1 Mode of Action Analysis of Liver Adenomas in Rats There are a number of possible mo des of action for hepatocarcinogene sis of chemicals. The weight of evidence of short-term genot oxicity assays on PFOA suggest that PFOA is not a DNA-reactive compound. As mitochondria play a major role in cell signali ng and apoptotic modes of cell death, and several structurally related perfluorinat ed compounds have been shown to manifest their toxicity by interfering with mitochondria biogenesis and bioenerge tics, the effects of PF OA on mitochondrial biogenesis and bioenergetics were investigated. A nu mber of studies have shown that mitochondria biogenesis in liver was increased following treatment of rats with PFOA ( e.g ., Berthiaume and Wallace, 2002). PFOA has also been demonstrated to uncouple oxidative phosphorylation in mitochondria of the liver from rats expos ed to PFOA in the diet(Keller et al., 1992). At high concentrations,PFOA caused a small increase in resting respiration rate and slightly decreases the membrane potential. The observed effects are be lieved to be attributed to a slight increas e in nonselective permeability of the mitochondria membranes caused by the surface-active property of the compound (Starkov and Wallace, 2002). Treatment with 200 μm PFOA was found to cause a dramatic increase in the cellular content of reac tive oxygen species (ROS) in human hepG2 cells. The activation of caspase-9 and apoptosis by PFOA observed was postula ted to be the result of the disruption of mi146OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

tochondria membrane and accumulation of ROS (Pan aretakis et al., 2001). Further research is needed, however, to elucidate how apoptosis is involved in tumori genesis of PFOA. Gap junctional intercellular communication (GJIC) , a process by which cells exchange ions, second messages, and other small molecules, is importantfor normal growth, development, and differentiation, as well as maintenance of homeostasis in muticellular organisms. Because tumor formation requires loss of homeostasis and many tumor promoters inhibit GJIC, it has been hy pothesized that GJIC may play a role in carcinogenesis (Trosko etal., 1998). PFOA has been demonstrated to inhibit GJIC in liver cells in vitro and in vivo (Upham et al., 1998). Since inhibition of GJIC is a widespread phenomenon, and the e ffect by PFOA was neither species nor tissue specific and was reversible, the significance of GJIC inhibition in regard to the mode of carci nogenic action of PFOA is unknown. Estrogen has been shown to promote hepatocarci nogenesis in rats (Yagerand Yager, 1980; Cameron et al., 1982). However, more research is need ed to support the involvement of this MOA in the hepatocarcinogenesis of PFOA. Low and sporadic incidences of liver necrosis were noted in both control and treated rats in the subchronic and chronic toxicity studies of PFOA. Liver necrosis and re generative cell proliferation may play a role in promoting pre-initi ated cells in carcinogenesis. The involvement of necrosis in the liver tumor induction by PFOA in rats remains to be investigated.While available data are not sufficient to support any of the above MOAs for the liver tumor induction by PFOA, there is strong evidence to conclude that the liver adenomas that are observed in rats following exposure to PFOA result from a PPAR α -agonist MOA. As descri bed in Klaunig et al. (2003), the MOA of PPAR α -agonist induced liver tumors involve s four causal key events which are shown in Figure 4 . The first key event is activation of PPAR α (which regulates the transcription of genes involved in peroxisome proliferation, cell cy cle control, apoptosis, and lipid metabolism). Activation of PPAR α leads to an increase in ce ll proliferation and a decrease in apoptosis, which in turn leads to preneoplastic cells and further cl onal expansion and formation of liver tumors. Of these key events, only PPAR α activation is highly specific for this MOA while cell proliferation/apoptosis and clonal expans ion are common to other modes of action. As depicted in Figure 4 , there are also several “associative” events that are markers of PPAR α agonism but are not directly involved in the etiology of liver tumors. These include peroxisome proliferation (a highly specific indicator that this MOA is operative) and peroxi somal gene expression. Peroxisomal proliferation may also result in hepatocyte oxidative stress wh ich may contribute to th e mode of action by causing indirect DNA damage and leading to mutations , or by stimulating cell proliferation. However, increases in oxidative damage to DNA have not been unambiguously demonstrated for PPAR α agonists. Oxidative stress is a general phenome non, and thus does not repr esent a highly specific marker for PPAR α -agonist induced liver carcinogenesis. Information that would help establish that a chemical is inducing li ver tumors via a PPAR α agonist MOA includes in vitro evidence of PPAR α agonism ( i.e ., evidence from an in vitro receptor assay), in vivo evidence of an increase in number and size of peroxisomes, increases in the activity of acyl CoA oxidase, and hepatic cell proliferation. The in vivo evidence should be collected from studies designed to provide the data needed to show dos e-response and temporal concordance between precursor events and liver tumor formation. Other info rmation that is desirable and may strengthen the weight of evidence for demonstrating that a PPAR α agonist MOA is operativ e includes data on hepatic CYP4A1 induction, palmitoyl CoA activity, he patocyte hypertrophy, incr ease in liver weights, decrease in the incidence of apoptosis, increase in microsomal fatty acid oxidation, and enhanced formation of hydrogen peroxide.

147OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Causative Events PPAR a Agonist

Activation of PPAR a

Cell Proliferation Decreased Apoptosis

Preneoplastic Foci

Clonal Expansion

Liver Tumors • Expression of Peroxisomal Genes

• Increase in Peroxisomes (number & size) Associative Events*

* Although there are other biological events (e.g., Kupffer cell mediated events, inhibition of gap junctions) , the measurements of peroxisome proliferation and peroxisomal enzyme activity (in particular acyl

- CoA ) are widely used as reliable markers of PPAR a activation.

Figure 4 Key events in the mode of action for PPAR α -agonis induced rodent liver tumors

There is sufficient information to demonstrate the key events for a PPAR α agonist MOA following exposure to PFOA in rodents. It has been well documented that PFOA is a potent peroxisome proliferator, inducing peroxisome proliferat ion in the liver of rats and mice ( e.g ., Ikeda et al., 1985; Pastoor et al., 1987; Sohlenius et al., 1992). Maloney and Waxman (1999) demonstrated that PFOA activates PPAR α using COS1 cells transfected with a lu ciferase reporter gene. Maximal transcriptional activity with PFOA was seen at 10

 M in the mouse PPAR α (and at 20

 M in human PPAR). Like many other peroxisome proliferators, PFOA ha s also been shown to cause hepatomegaly (an early biomarker of peroxisome proliferator hepato carcinogenesis) in rats (T akagi et al., 1992; Cook, 1994) and mice (Kennedy, 1986), and induc e oxidative DNA damage in li ver of rats (Takagi et al., 1991).
Several studies have been conducted to examine the dose-response and te mporal relationships among key endpoints. The temporal and dose-respons e relationship of measur es of peroxisome proliferation, hepatocellular hype rtrophy, liver weight, and liver hist opathology have been examined in male Sprague-Dawley rats following 4, 7 and 13 weeks of administration of dietary PFOA at doses ranging from 0-6.5 mg/kg-day (Palazzolo, 1993). The results are summarized in Table 46. There was no evidence of peroxisome proliferation, hepa tocellular hypertrophy or liver weight increases at 0.06 mg/kg-day. At doses rangi ng from 0.64 to 6.5 mg/kg-day th ere is a clear relationship between peroxisome proliferation (indicated by incr eased palmitoyl CoA oxid ase activity), hepatocellular hypertrophy and increases in liv er weight at all time points. There was no evidence of hepatocellular necrosis.

148OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 46 Summary of Liver Effects in Male Sprague-Dawley Rats Fed APFO for 90 Days
Dose (mg/kg-day) 1 Parameter Week

0 a 0 b 0.06 0.64 1.94 6.5 Palmitoyl CoA Oxidase (IU/G) 4
7
13 8 (0.5) 7 (1.5) 8 (0.9) 5 (0.4) 7 (1.5) 5 (1.1) 9 (1.7) 7 (0.8) 8 (1.9) 14 (3.8) c 18 (5.5) 10 (2.1) 24 (11.4) c 32 (12.2) c 14 (3.4) c 37 (14.8) cd 54 (35.3) cf 17 (4.5) cd Hepatocellular Hypertrophy 4
7
13 0/15
0/15
0/15 0/15
0/15
0/15 0/15
0/15
0/15 12/15
12/15
13/15 15/15
15/15
14/15 14/15
15/15
15/15 Hepatocellular Necrosis, Coagulative 4
7
13 0/15
0/15
0/15 1/15
0/15
0/15 0/15
0/15
1/15 0/15
0/15
0/15 1/15
0/15
1/15 2/15
1/15
0/15
Absolute Liver Weight (g) 4

7

13 16.34 (2.14)
17.78 (2.12)
19.73 (2.01) 15.83 (1.13)
16.91 (2.22)
16.30 (1.62) 15.45 (1.71)
17.68 (??)
18.03 (2.81) 17.89 (2.13)
19.42 (2.10)
20.44 (2.87) 23.23 (2.83) c
27.76 (3.51) c
22.74 (4.21) 25.44 (1.89) cd

27.76 (3.51) cd

26.78 (5.47) cd Liver/Body Weight (%) 4

7

13 3.97 (0.37)
3.75 (0.29)
3.53 (0.28) 4.07 (0.27)
3.76 (0.37)
3.23 (0.23) 3.73 (0.23)
3.64 (0.33)
3.24 (0.30) c 4.48 (0.32) d

4.12 (0.37)
3.69 (0.32) 5.77 (0.60) d

5.14 (0.53) c
4.21 (0.56) c 6.73 (0.49) de

6.06 (0.59) cd

5.49 (0.84) cd
1- Mean (SD) a - non-pair-fed controls b - pair-fed controls c - statistically significant at p < 0.05 with the non-pair-fed control d - statistically significant at p < 0.05 with the pair-fed control e - calculated using th e non-pair-fed control f - calculated using the pair-fed control Liu et al. (1996) characterized th e dose-response relationships of se veral key endpoints in male CD rats exposed to doses of 0.2, 2, 20, and 40 mg/kg-day PFOA for 14 days. These endp oints included liver weight, hepatic β -oxidation, hepatic aromatase (P450 19A 1), and hepatic total cytochrome P450. The NOEL for these endpoints was found to be 0.2 mg/kg-day with significant changes observed at > 2 mg/kg-day for all endpoints. Thus, the studies of Palazzolo (1993) and Liu et al. (1996) demonstrate evidence of peroxisome proliferation at doses close to and below doses that result in liver adenomas fo llowing chronic exposures (1.5 and 15 mg/kg-day). Biegel et al. (2001) examined the temporal rela tionship between relative liver weights, hepatic

 β

oxidation, and hepatic cell proliferation, and hepatic adenomas were evaluated in CD rats following PFOA exposure for 1, 3, 6, 9, 12, 15, 18, 21, and 24 months. Relative liver weights and hepatic

 β

oxidation were increased at all time points. Hepatic cell proliferation was numerically increased relative to the pair-fed control at 9, 15, 18, and 21 mont hs. The liver end points (weight,

 β -oxidation,

149OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

and cell proliferation) were all el evated well before the first occurrence of liver adenomas, which occurred after 12 months of treatment. Significant increases in liver weig ht were also observed in wild -type mice exposed to PFOA, or Wy-14,643 (a benchmark peroxisome proliferator) intheir diet at 0.02% and 0.125% (w/w), respectively, for 7 days (Yang et al., 2002a). This re sponse was totally absent upon treatment of PPAR α

null mice with Wy-14,643, but, surprisi ngly, virtually unaltered in nu ll mice exposed to PFOA. The finding appears to suggest that hepatomegalymay be induced by PFO A independent of PPAR α in mice. Since this is only a short-te rm study in mice, further research is required to de monstrate if the hepatomegaly induced by PFOA in rat is PPAR α -independent.
Conclusions The data clearly demonstrate that PFOA indu ces liver toxicity and adenomas via a PPAR α agonist MOA in rats. PFOA activates the PPAR α and the requisite dose-respon se and/or temporal associations of the key events for the PPAR α mode of action with the liver adenomas have been characterized.

3.7 Prenatal Developmental Toxicity Studies in Animals Several prenatal developmental toxicity studies of APFO have been conducted. These include two oral studies in rats, one oral study in rabbits, one oral study in mice, and one inhalation study in rats. Gortner (1981) administered time-mated Sprague-Dawley rats (22 per group) doses of 0, 0.05, 1.5, 5, and 150 mg/kg-day APFO in distilled water by ga vage on gestation days (GD) 6-15. Doses were adjusted according to body weight . Dams were monitored on GD 3-20 for clinical signs of toxicity. Individual body weights were recorded on GD 3, 6, 9, 12, 15, and 20. Animals were sacrificed on GD 20 by cervical dislocation and th e ovaries, uteri, and contents we re examined for the number of corpora lutea, number of viable and non-viable fetuses, number of resorption sites, and number of implantation sites. Fetuses were weighed and sexe d and subjected to external gross necropsy. Approximately one-third of the fetuses were fixed in Bouin’s solution and exam ined for visceral abnormalities by free-hand sectioning. The remaining fetuses were subjected to skeletal examination using alizarin red. Signs of maternal toxicity consisted of statistically significant reductions in mean maternal body weights on GD 9, 12, and 15 at the high-dose group of 150 mg/kg-day. Mean maternal body weight on GD 20 continued to remain lower than controls, although the difference wa s not statis tically significant. Other signs of maternal toxicity that occurred only at the high-dos e group included ataxia and death in three rat dams. No other effects we re reported. Administrati on of APFO during gestation did not appear to affect the ovaries or reproductive tract of the dams. Under the conditions of the study, a NOAEL of 5 mg/kg-day and a LOAEL of 150 mg/kg-day for mate rnal toxicity were indicated. A significantly higher incidence infetuses with one missing sterne brae was observed at the highdose group of 150 mg/kg-day; however this skeletal variation also oc curred in the controls and the other three dose groups (at similar incidence but lower than the high-dose group) and therefore was not considered to be treatment-related. No signi ficant differences between treated and control groups were noted for other developmental paramete rs that included the meannumber of males and females, total and dead fetuses, the mean number of resorption sites, implantation sites, corpora lutea and mean fetus weights. Like wise, a fetal lens finding initiallydescribed as a variety of abnormal morphological changes localized to the area of the embryonal nucleus, was later determined to be an artifact of the free-hand sectioning technique and therefore not cons idered to be treatment150OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

related. Under the conditions of the study, a NOA EL for developmental toxi city of 150 mg/kg-day (highest dose group) was indicated. Staples et al. (1984) also conducte d a developmental toxicity study ofAPFO in rats. The study design consisted of an inhalation and an oral portion, eachwith two trials or e xperiments. In the first trial the dams were sacrificed on GD 21; while in the second trial, the dams were allowed to litter and the pups were sacrificed on day 35-post partum. For the inhalation portion of the study, the two trials consisted of 12 pregnant Sprague-Dawle y rats per group exposed to 0, 0.1, 1, 10, and 25 mg/m 3 APFO by whole-body dust inhalati on for 6 hours/day, on GD 6-15. In the oral portion of the study, 25 and 12 Sprague-Dawley rats for the first and second tria ls, respectively, were administered 0 and 100 mg/kg-day APFO in corn oil by ga vage on GD 6-15. For both routes of administration, females were mated on an as-needed basis a nd when the number of mated females was bred, they were ranked within breeding days by body weight and assi gned to groups by rotation in orderof rank. Finally, two additional groups (six dams per group) were added to each trial that was pairfed to the 10 and 25 mg/m 3 groups. For trial one, the dams were weighed on GD 1, 6, 9, 13, 16, and 21 and observed daily for abnormal clinical signs. On GD 21, the dams were sacrif iced by cervical dislocation and examined for any gross abnormalities, liver weights were recorded and the reproductive status of each animal was evaluated. The ovaries, uterus and co ntents were examined for the num ber of corpora lutea, live and dead fetuses, resorptions and implantation sites. Pups (live and dead) were counted, weighed and sexed and examined for external, visceral, and sk eletal alterations. The heads of all control and high-dosed group fetuses were examined for visceral alterations as well as macro- and microscopic evaluation of the eyes. For trial two, in which the dams were allowed to li tter, the procedure was the same as that for trial one up to GD 21. Two days before the expected day of parturition, each dam was housed in an individual cage. The date of partur ition was noted and designated Da y 1 PP. Dams were weighed and examined for clinical signs on Days 1, 7, 14, and 22 PP. On Day 23 PP all dams were sacrificed. Pups were counted, weighed, and examined for ex ternal alterations. Eachpup was subsequently weighed and inspected for adverse clinical signs on Days 4, 7, 14, and 22 PP. The eyes of the pups were also examined on Days 15 and 17 PP for th e inhalation portion and on Days 27 and 31 PP for the gavage portion of the study. Pups were sacrificed on Day 35 PP and examined for visceral and skeletal alterations. In trial one of the inhalation study, treatment-related clinical signs of maternal toxic ity occurred at 10 and 25 mg/m 3 and consisted of wet abdomens, chromodacryorrhea, chromorhinorrhea, a general unkempt appearance, and lethargy in four dams at the end of the exposure period (highconcentration group only). Three out of 12 dams died during treatment at 25 mg/m 3 (on GD 12, 13, and 17). Food consumption was signifi cantly reduced at both 10 and 25 mg/m 3 ; however, no significant differences were noted between treated and pair-fed groups. Signi ficant reductions in body weight were also observed at these concentrati ons, with statistical significance at the highconcentration only. Likewise, statis tically significant increases in mean liver weights (p < 0.05) were seen at the high-concentration group. Under the conditions of the study, a NOAEL and LOAEL for maternal toxicity of 1 and 10 mg/m 3 , respectively, were indicated. No effects were observed on the maintenance of pr egnancy or the inciden ce of resorptions. Mean fetal body weights were signifi cantly decreased in the 25 mg/m 3 groups (p = 0.002) and in the control group pair-fed 25 mg/m 3 (p = 0.001). However, interpretation of the decreased fetal body weight is difficult given the high incidence of mo rtality in the dams. Under the conditions of the study, a NOAEL and LOAEL for developm ental toxicity of 10 and 25 mg/m 3 , respectively, were indicated.

151OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

In trial two of the inhalation st udy, clinical signs of maternal toxicity seen at 10 and 25 mg/m 3 were similar in type and inci dence to those described for trial one. Maternal body weight gain during treatment at 25 mg/m 3 was less than controls, although the di fference was not statistically significant. In addition, 2 out of 12 dams died during treatment at 25 mg/m 3 . No other treatment-related effects were reported, nor were any adverse effects noted for any ofthe measurements of reproductive performance. Under the condi tions of the study, a NOAEL and LO AEL for maternal toxicity of 1 and 10 mg/m 3 , respectively, were indicated. Signs of developmental toxicity in this group consis ted of statistically signif icant reductions in pup body weight on Day 1 PP (6.1 g at 25 mg/m 3 vs. 6.8 g in controls). OnDays 4 and 22 PP, pup body weights continued to remain lower than controls , although the difference wa s not statistically significant (Day 4 PP: 9.7 g at 25 mg/m 3 vs.10.3 in controls; Day 22 PP: 49.0 g at 25 mg/m 3 vs. 50.1 in controls). No significant effects were reported fo llowing external examinati on of the pups or with ophthalmoscopic examination of the eyes. Again, inte rpretation of these effects is problematic given the high incidence of maternal mortality. Under the conditions of the study, a NOAEL and LOAEL for developmental toxicity of 10 and 25 mg/m 3 , respectively, were indicated. In trial one of the oral study, three out of 25 da ms died during treatment of 100 mg/kg APFO during gestation (one death on GD 11; two on GD 12). Clinical signs of matern al toxicity in the dams that died were similar to those seen with inhalation exposure. Food consumption and body weights were reduced in treated animals compared to controls. No adverse signs of toxicity were noted for any of the reproductive parameters such as maintenance of pregnancy or incidence of resorptions. Likewise, no significant differences between treated and control groups were noted for fetal weights, or in the incidences of malformati ons and variations; nor were ther e any effects noted following microscopic examination of the eyes. In trial two of the oral study, sim ilar observations for clinical signs were noted for the dams as in trial one. Likewise, no adverse effects on reproductiv e performance or in any of the fetal observations were noted. An oral prenatal developmental toxicity study wa s conducted in rabbits (G ortner, 1982). Based on the results of a range-finding study, an upper dose level of 50 mg/kg- day was set for the definitive study in which four groups of 18 pregnant New Zealand White rabb its were administered 0, 1.5, 5, and 50 mg/kg-day APFO in distilled water by gava ge on gestation days 6-18. Pregnancy was established in each sexually mature female by i.v. injec tion of pituitary lutenizi ng hormone in order to induce ovulation, followed by artificial insemination with 0.5 ml of pooled semen collected from male rabbits; the day of insemination was designat ed as day 0 of gestati on. A constant dose volume of 1 ml/kg was administered. Individual body weights were measured on GD 3, 6, 9, 12, 15, 18, and 29. The does were observed daily on GD 3-29 for a bnormal clinical signs. On GD 29, the does were euthanized and the ovaries, uterus and contents ex amined for the number of corpora lutea, live and dead fetuses, resorptions and implantation sites. Fetuses were examined for gross abnormalities and placed in a 37 0 C incubator for a 24-hour survival check. Pups were subsequently euthanized and examined for visceral and skeletal abnormalities. A blood sample was taken from six does prior to dosing and then on GD 18 and 29; a liver sample was taken from the same animals on GD 29. All samples were sent to the sponsor for analysis. Th is information was unavailable at the time of this review. Signs of maternal toxicity cons isted of statistically significant transient reductions in body weight gain on GD 6-9 when compared to controls; body weight gains return ed to control levels on GD 12- 29. Administration of APFO during ge station did not appear to aff ect the ovaries or reproductive tract contents of the does. No clinical or other treatment-related signs were reported. Under the conditions of the study, a NOAEL of 50 mg/kg-day, the highest dose tested, for maternal toxicity was indicated.

152OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

No significant differences were not ed between controls and treated groups for the number of male and female fetuses, dead or live fetuses, or fetal weights. Likewise, there we re no significant differences reported for the number of resorption and implantation sites, corpora lutea, the conception incidence, abortion rate, or th e 24-hour mortality incidence of the fetuses. Gross necropsy and skeletal/visceral examinations were unremarkable. The only sign of deve lopmental toxicity consisted of a dose-related increase in a skeletal va riation, extra ribs or 13 thrib, with statistical significance at the high-and mid-dose gr oups (38% at 50 mg/kg/day, 30% at 5 mg/kg/day, 20% at 1.5 mg/kg/day, and 16 % at 0 mg/ kg/day). The incidence of 13 thrib at the low-dose group occurred at a slightly higher rate, but was not statistically significantly different from controls. Therefore, under the conditions of the study, a L OAEL for developmental toxicity of 5 mg/kg-day was indicated. Lau et al. (2005) recently conducted a developmenta l toxicity study of APFO in mice in order to ascertain whether there was a sex difference in the bioaccumulation of APFO in the mouse, and to evaluate the effects of APFO on pre- and postnatal development inoffspring exposed during pregnancy. In that study, groups aver aging 9-45 timed-pregnant CD-1 mice were given 0, 1, 3, 5, 10, 20, or 40 mg/kg APFO daily by oral gavage on GD 1- 17. Maternal weight was monitored during gestation. Dams were divided into two groups. In the fi rst group, dams were sacrificed on GD 18 and underwent maternal and fetal examin ations that included measure of maternal liver weight, and examination of the gravid uterus for numbers of liv e and dead fetuses and re sorptions. Maternal blood was collected and analyzed for APFO serum concentr ation. APFO levels in the fetuses were not examined. Live fetuses were weighed and subjected to external gross necropsy and skeletal and visceral examinations. In the second group of dams, an additional dose of AP FO was administered on GD 18. Dams were allowed to give birth on GD 19. The day following partur ition was designated as postnatal day (PND) 1. Time of parturition, condition of newborns, and number of live offspring were recorded. The number of live pups in each lit ter and pup body weight were noted for the first four days after birth and then at corresponding in tervals thereafter. Eye opening was recorded beginning at PND 12. Pups were weaned on PND 23 and separated by sex. The time to sexual maturity was determined by monitori ng vaginal opening and preputia l separation beginning on PND 24. Two to four pups per sex per litt er were randomly selected for obs ervation of postnatal survival, growth, and development. Estrous cyclicity was determined daily by vaginal cytology. After weaning, dams were sacrificed and the contents of the uteri examined fo r implantation sites. Postnatal survival was calculated based on the number of imp lantations for each dam. Benchmark dose values (BMD 5 and BMDL 5 ) were calculated for the maternal and developmental toxic endpoints. In order to compare the body burdens of APFO betw een rat and mouse after subchronic exposure, a separate study was conducted in which 0 mg/kg/day, 10 mg/kg (for 20 days) and 20 mg/kg (for 17 days) APFO was administered by gavage to adu lt male and female rats and mice. Twenty-four hours after the last treatment, the animals were sacrificed and blood samples collected for subsequent APFO serum analyses.
Serum concentrations of PFOA for controls and tr eated animals were analyzed by high-performance liquid chromatography with mass sp ectrometry by spiking aliquots of thawed serum sample with 13 C-PFOA internal standard. Signs of maternal toxicity were observed follo wing exposure to APFO during pregnancy. Slight, but statistically significant (p < 0.05), increases in the average time to parturition were observed at 10 and 20 mg/kg/day. Dose-related decreases in ma ternal weight gain dur ing pregnancy were observed beginning at 5 mg/kg/day, w ith statistical significance (p < 0.05) seen in the 20 and 40 mg/kg/day dose groups. Sta tistically significant (p < 0.05) dose-related increas es in maternal liver weight were also observed, be ginning at 1 mg/kg/day. Signs of developmental toxicity were observed fo llowing in utero exposure to APFO. Statistically significant (p < 0.05) increases in full- litter resorptions were reported at doses of > 5 mg/kg/day,

153OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

with complete loss of pregnancies at the highe st dose group of 40 mg/kg/day. However, the number of implantations in treated groups was comparable to controls, which according to the authors, may indicate the pregnancy losses may have occurred s hortly after implantation. In litters with live fetuses at term, statistically significant (p < 0.05) increases in prenatal loss were observed in the 20 mg/kg/day dose group. Increases (p < 0.05) in stillbirths and neonata l mortality (or decreases in postnatal survival) were observed at doses > 5 mg/kg/day, with as much as a 30% increase in these effects seen at 10 and 20 mg/kg/day dose group. Po stnatal survival and vi ability at the 1 and 3 mg/kg/day dose groups was comparable to controls. Statistically significant (p < 0.05) increases in delayed or reduced ossification of various bones were observed at doses of 10 and 20 mg/kg/day. Statistically significant (p < 0.05) increases in minor limb and tail defects were observed in the fetuses at doses > 5 mg/kg/day. A 20% reduction (p < 0.05) in live fetal body weight at term was reported at 20 mg/kg/day. At doses > 3mg/kg/day, a trend in growth retardation (p < 0.05), as manifested by reductions (25-30%) in body weights, was observed in the neonates at weaning. Body weights returned to control levels by 6 weeks of age for females, and by 13 weeks of age for males, with body weights exceeding those of controls by 10% as the animals continued to age. Deficits in early postnatal growth and development we re also manifested by significant (p < 0.05) delays in eye opening at doses > 5 mg/kg/day. Slight delays (p < 0.05) in vaginal opening and in time to estrous were observed at 20mg/kg/day in females; which was in contrast to significant accelerations (p < 0.05) in sexual maturation observed in males, with preputial separation o ccurring 4 days earlier than controls at the 1 mg/kg/day; and 2-3 days ea rlier in the 3 through 10 dos e groups; but not at the 20 mg/kg/day dose group, which was slightlydelayed compared to controls. Serum analyses in adult male and female rats fo llowing subchronic exposure s indicated differences in the accumulation of APFO. Fo llowing subchronic exposures of 20 days, serum concentration of APFO in male rats exposed to 10 mg/kg APFO was 111 μg/ml, while only 0.69 μg/ml was detected in female rats. This is in contrast to what was ob served in adult mice; no significant differences in serum APFO concentrations were observed between the genders. In fact, a steady-state in APFO serum concentration in mice was reached by 7 days and remained fairly constant for the rest of the exposure period of 17 days. Serum levels in mice exposed to 20 mg/kg APFO recorded at 7 and 17 days of treatment were the same for both males and females, approximately 180-200 μg/ml. Serum analysis for pregnant mice revealed a dose-depe ndent increase in accumulation of APFO at term. Based on the results observed in the st udy, the authors calculated various benchmark dose values for maternal and developm ental toxic endpoints: For maternal toxicity, BMD 5 and BMDL 5 estimates for decreases in maternal weight gain during pregnancy were 6.76 and 3.58 mg/kg, respectively. For increases in maternal liver weight at term, BMD 5 and BMDL 5 estimates were 0.20 mg/kg and 0.17 mg/kg, respectively. For developmental toxicity, BMD 5 and BMDL 5 estimates for the incidence of full-litter resorptions and neonatal mortality (determined by survival to weaning) observed at the 5 mg/kg/day dose group were 2.84 and 1.09 mg/kg, respectively. Significant alterations in postn atal growth and development were observed at 1 and 3 mg/kg/day, with BMD 5 and BMDL 5 estimates of 1.07 and 0.86 mg/kg, respectively, for decreased pup weight at wean ing; and 2.64 and 2.10 mg/kg, respectively, for delays in eye opening. The BMD 5 and BMDL 5 estimates for reduced phalang eal ossification were < 1 mg/kg. BMD 5 and BMDL 5 estimates for reduced fetal weight at term were estimated to be 10.3 and 4.3 mg/kg, respectively. According to the study authors, the results of th e current study in mice suggest significant differences in the developmental toxicity profile than th at previously reported in rats. The authors concluded that these differences in toxicity could be due, in part, to the differential pharmacokinetic handling and disposition of AP FO between rats and mice.

154OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Conclusions In an oral prenatal developmental toxicity study in rats, the LOAEL and NOAEL for maternal toxicity were 150 mg/kg-day and 5 mg/kg-day, respectively. There was no evidence of developmental toxicity after exposure to doses as high as 150 mg/kg-day. In a rat inhalation developmental toxicity study, the NOAEL and LOAEL for maternal toxicity were 1 and 10 mg/m 3 , respectively. The NOAEL and LOAEL for developmental to xicity were of 10 and 25 mg/m 3 , respectively. In a rabbit oral prenatal developmental toxic ity study there was a significant incr ease in skeletal variations after exposure to 50 mg/kg-day APFO. There was no evid ence of maternal toxi city at 50 mg/kg-day, the highest dose tested. In a mous e oral developmental toxicity study, there was evidence of maternal toxicity (decreased weight gain and increased liver weight; the BMD 5 and BMDL 5 estimates for decreases in maternal weight gain were 6.76 and 3.58 mg/kg, respectively, and the BMD 5 and BMDL 5 estimates for increases in ma ternal liver weight were 0.20 mg/kg and 0.17 mg/kg, respectively. There was also evidence of developmental t oxicity in the mouse study and the authors calculated benchmark doses for a variety of endpoints. The BMD 5 and BMDL 5 estimates for the incidence of full-litter resorptions and neonatal mortality (determined by survival to weaning) observed at the 5 mg/kg/day dose group were 2.84 and 1.09 mg /kg, respectively. Signi ficant alterations in postnatal growth and development were observed at 1 and 3 mg/kg/day, with BMD 5 and BMDL 5 estimates of 1.07 and 0.86 mg/kg, respectively, for decreased pup weight at weaning; and 2.64 and 2.10 mg/kg, respectively, for de lays in eye opening. The BMD 5 and BMDL 5 estimates for reduced phalangeal ossification were < 1 mg/kg. BMD 5 and BMDL 5 estimates for reduced fetal weight at term were estimated to be 10.3 and 4.3 mg/kg, respectively.

3.8 Reproductive Toxicity Studies in Animals An oral two-generation reproductiv e toxicity study of APFO in rats was conducted (York, 2002; Butenhoff et al., 2004a). Five groups of 30 Sprague -Dawley rats per sex per dose group were administered APFO by gavage at dos es of 0, 1, 3, 10, and 30 mg/kg-day six weeks prior to and during mating. Treatment of the F0 male rats continued until mating was confirmed, and treatment of the F0 female rats continued throughout ge station, parturiti on, and lactation. The F0 animals were examined twice daily for clin ical signs, abortions, pr emature deliveries, and deaths. Body weights of F0 male rats were record ed weekly during the dosage period and then on the day of sacrifice. Body weights of F0 female ra ts were recorded weekly during the pre- and cohabitation periods and then on gestation days (GD) 0, 7, 10, 14, 18, 21, and 25 (if necessary) and on lactation days (LD) 1, 5, 8, 11, 15, and 22 (termina l body weight). Food consumption values in F0 male rats were recorded weekly during the treatment period, while in F0 female rats, values were recorded weekly during the precohabitation period, on GDs 0, 7, 10, 14, 18, 21, and 25 and on LDs 1, 5, 8, 11, and 15. Estrous cycling was evaluated daily by examinati on of vaginal cytology beginning 21 days before the scheduled cohabitation period and continuing until confirmation of mating by the presence of sperm in a vaginal smear or confirmation of a c opulatory plug. On the day of scheduled sacrifice, the stage of the estrous cycle was assessed. For mating, one male rat and one female rat per group were cohabitated for a maximum of 14 days. Female rats with evidence of sperm in a vaginal smear or copulatory plug were designated as GD 0. Parental females were evaluated for length of gest ation, fertility index, gest ation index, number and sex of offspring per litter, number of implantation sites, general condition of the dam and litter during the postpartum period, litter size and viability, vi ability index, lactation i ndex, percent survival, and sex ratio. Maternal beha vior of the dams was reco rded on LDs 1, 5, 8, 15, and 22.

155OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

F0 generation animals were sacrificed by carbon di oxide asphyxiation (day 106 to 110 of the study for male rats, i.e., after completion of the cohabitation period; and LD 22 for female rats), necropsied, and examined for gross lesions. Gross necropsy included examination of external surfaces and orifices, as well as internal exam ination of tissues and organs. Indi vidual organs were weighed and organ-to-body weight and organ-to -brain weight ratios were cal culated for the brain, kidneys, spleen, ovaries, testes, thymus, liver, adrenal glands , pituitary, uterus with oviducts and cervix, left epididymis (whole and cauda), right epididymis, pr ostate and seminal vesicles, (with coagulating glands and with and without flui d). Tissues retained in neutral buffered 10% formalin for possible histological evaluation included th e pituitary, adrenal glands, vagina , uterus, with oviducts, cervix and ovaries, right testis, seminal vesicles, right epididymis, and prostate. Histological examination was performed on tissues from 10 randomly selected rats per sex from the control and high dosage groups. All gross lesions were examined histologicall y. All F0 generation rats that died or appeared moribund were also examined. Histological examination of the reproductive organs in the low- and mid-dose groups was conducted in rats that exhibited reduced fertility by either failing to mate, c onceive, sire, or deliver healthy offspring; or for which estrous cyclic ity or sperm number, mo tility, or morphology were altered. Sperm number, motility, and morphology were ev aluated in the left cauda epididymis of F0 generation male rats; testicular spermatid concentrations were evaluated in the left testis. The number and distribution of implantation sites were recorded in F0 generation female rats. Rats that did not deliver a litter were sacrific ed on GD 25 and examined for pregna ncy status. Uteri of apparently nonpregnant rats were examined to confirm the ab sence of implantation sites. A gross necropsy of the thoracic, abdominal and pelvic viscera was pe rformed. Female rats without a confirmed mating date that did not deliver a li tter were sacrificed on an es timated day 25 of gestation. At scheduled sacrifice, after completion of the cohabitation period in F0 male rats and on LD 22 in F0 female rats, blood samples (10 males and 10 females each for the 10 and 30 mg/kg-day dose groups; 3 males and 3 females for the control gro up) were collected for analysis of PFOA. The F1 generation pups in each litter were counted once daily. The litter sizes were not standardized on day 4 or at anytime during lactat ion. Physical signs (including vari ations from expected lactation behavior and gross external physical anomalies) were recorded for the pups each day. Pup body weights were recorded on LDs 1, 5, 8, 15 and 22. OnLD 12, all F1 generation male pups were examined for the presence of nipples. Pups that died before examination of the litter for pup viability on LD 1 were evaluated for vital status at birth. Pups found dead on LDs 2 to 22 were examined for gross lesions and for the cause of death. All F1 generation rats were weaned on LD 22 based on observed growth and viability of these pups. At weaning (LD 22), two F1 generation pups per se x per litter per group (60 male and 60 female pups per group) were selected for continued evaluati on, resulting in 600 total rats (300 rats per sex) assigned to the five dosage groups . At least two male pups and tw o female pups per litter, when possible, were selected. F1 generation pups not se lected for continued observation for sexual maturation were sacrificed. Three pups per sex per litte r were examined for gro ss lesions. Necropsy included a single cross-section of the head at the le vel of the frontal-parietal suture and examination of the cross-sectioned brain for apparent hydroce phaly. The brain, spleen and thymus from one of the three selected pups per sex pe r litter were weighed and the br ain, spleen, and thymus from the three selected pups per sex per litter were retain ed for possible histological evaluation. All remaining pups were discarded without further examination. The F1 generation rats were given the same dosa ge level of the test substance and in the same manner as their respective F0 generation sires and dams. Dosages were gi ven once daily, beginning at weaning and continuing until the day before sacrifice. F1 generation female rats were examined for age of vaginal patency, beginning on day 28 pos tpartum (LD 28). F1 generation male rats were

156OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

evaluated for age of preputial separation, begi nning on day 39 postpartum (LD 39). Body weights were recorded when rats reached sexual maturation. Following sexual maturation, a table of random units was used to select one male and one female per litter per group for continuation through mating to produce the F2 generation. The remaining F1 animals were sacrificed. Estrous cycling was evaluated daily by examinati on of vaginal cytology beginning 21 days before the scheduled cohabitation period and continuing until confirmation of mating by the presence of sperm in a vaginal smear or confirmation of a c opulatory plug. On the day of scheduled sacrifice, the stage of the estrous cycle was assessed. A table of random units was used to assign F1 gene ration rats to cohabitation, one male rat per female rat. If random assignment to cohabitation resulted in the pair ing of F1 generation siblings, an alternate assignment was made. The cohabitation period consisted of a maximum of 14 days. Body weights of the F1 generation male rats were recorded weekly during the postweaning period and on the day of sacrifice. Body weights of the F1 generation fema le rats were recorded weekly during the postweaning period to cohabitation, and on GDs 0, 7, 10, 14, 18, 21 and 25 (if necessary) and on LDs 1, 5, 8, 11, 15 and 22. Food consumption valu es for the F1 generation male rats were recorded weekly during the dosag e period. Food consumption values for the F1 generation female rats were recorded weekly during the postwean ing period to cohabitation, on GDs 0, 7, 10, 14, 18, 21 and 25 and on LDs 1, 5, 8, 11 and 15. Because pups begin to consume maternal food on or about LD 15, food consumption values were not tabulated after LD 15. At scheduled sacrifice, the F1 animals were subject ed to gross necropsy, a nd selected organs were weighed and examined histologically as described above for the F0 animals. Sperm analyses were also conducted as described for the F0 animals. F2 generation litters were examined after deliver y to identify the numberand sex of pups, stillbirths, live births and gross alterations. Each litter was evaluated for viability at least twice each day of the 22-day postpartum period. D ead pups observed at these times were removed from the nesting box. Anogenital distance was measured for all live F2 generation pups on LDs 1 and 22, and F2 male pups were examined for the presence of nipples on LD 12.

3.8.1 F0 Generation

3.8.1.1 F0 Males One F0 male rat in the 30 mg/kg-day dose group wa s sacrificed on day 45 of the study due to adverse clinical signs (emaciation, cold-to-touch, and decreased motor activity). Necroscopic examination in that animal revealed a pale and tan liver and red testes. All other F0 generation male rats survived to scheduled sacrifice. Statistically sign ificant increases in clinical signs were also observed in male rats in the high-dose group that included dehydration, urine- stained abdominal fur, and ungroomed coat. Significant reductions in body weightand body weight gain were repo rted for most of the dosage period and continuing until term ination of the study in the 3, 10, and 30 mg/kg-day dose groups. Absolute food consumption values were also significantly reduced during these periods in the 30mg/kg-day dose group, while significant increases in relative food consumption values were observed in the 3, 10, and 30 mg/kg-day within those same periods.

157OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

No treatment-related effects were reported at any dose level for any of the mating and fertility parameters assessed, including numbers of days to in seminate, numbers of rats that mated, fertility index, numbers of rats with conf irmed mating dates during the firs t and second week of cohabitation, and numbers of pregnant rats per rats in co habitation. At necropsy, none of the sperm parameters evaluated (sperm number, motility, or morphol ogy) were affected by treatment at any dose level. At necropsy, statistically significa nt reductions in terminal body we ights were seen at 3, 10, and 30 mg/kg-day (6%, 11%, and 25% decreas e from controls, respectively; p < 0.05). Absolute weights of the left and right epididymides, left cauda epididymis, seminal ve sicles (with and without fluid), prostate, pituitary, left and right adrenals, spleen, and thymus were also statistically significantly reduced at 30 mg/kg-day. These absolute organ weight reductions are probably due to reductions in body weight and not a reflection of target organ toxicity since the organ-to-body weight ratios were either normal or significantly in creased. The biological significance of the weight changes for the adrenals, however, is unclear since dose-related histopathological changes were observed. The absolute weight of the seminal vesicles without fl uid was significantly redu ced in the 10 mg/kg-day dose group. The absolute weight of the liver was statistically significantlyincreased in all dosegroups. Absolute kidney weights we re statistically significantly in creased in the 1, 3, and 10 mg/kgday dose groups, but significantly decreased in th e 30 mg/kg-day group. Organ weight-to-terminal body weight ratios for the liver an d for the left and right kidney we re statistically significantly increased in all treated groups. The biological signifi cance of the weight change s observed in the liver and kidney is unknown since no histopathology was conducted on these organs. Organ weight-toterminal body weight ratios for th e brain were statistically signifi cantly increased for the 3, 10, and 30 mg/kg-day dose groups; organ weight-to-brain wei ght ratios were significantly reduced for some organs at the high dose group, and si gnificantly increased for other organs among all treated groups. No treatment-related effects were seen at necrop sy or upon microscopic examination of the reproductive organs, with the exception of increased th ickness and prominence of the zona glomerulosa and vacuolation of the cells of the adrenal cort ex in 2/10 males in the 10 mg/kg-day group and 7/10 males in the 30 mg/kg-day dose group. Serum analysis for the F0 generation males in the control, 10 and 30 mg/kg-day groups sampled at the end of cohabitation showed that PFOA was pres ent in all samples tested, including controls. Control males had an average concentration of 0.0344 + 0.0148 μg/ml PFOA. Levels of PFOA were similar in the two male dose groups; treated males had 51.1 + 9.30 and 45.3 + 12.6 μg/ml, respectively for the 10 and 30 mg/kg-day dose groups.

3.8.1.2 F0 Females

No treatment-related deaths or adverse clinical si gns were reported in parental females at any dose level. No treatment-related effects were reported for body weights, body weight gains, and absolute and relative food c onsumption values. There were no treatment-related effects on estrous cy clicity, mating or fertility parameters. None of the natural delivery and litter observations were a ffected by treatment, that is, the numbers of dams delivering litters, the duration of gestation, the averages for implan tation sites per delivered litter, the gestation index (number of dams with one or more liveborn pups/number of pregnant rats), the numbers of dams with stillborn pups, dams with a ll pups dying, liveborn and stillborn pups viability index, pup sex ratios, and mean birth weights were comparable to controls among all treated groups. Necropsy and histopathol ogical evaluation were al so unremarkable. Termin al body weights, organ weights, and organ-to-terminal body weight ratios were comparable to control values for all treated

158OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

groups, except for kidney and liver weights. The absolute weights of the left and ri ght kidney, and the ratios of these organ weights-to-terminal b ody weight and of the leftkidney weight-to-brain weight were significantly reduced at the highest dose of 30 mg/kg- day. The biological significance of these weight changes is not known since histopathology was not conducted on the kidney. The ratio of liver weight-to-terminal body weight was significantly reduced at3 and 10 mg/kg-day, but there were no effects observed at 30 mg/kg-day or in absolute liver weight. Serum PFOA levels were analyzed from the control, 10 and 30 mg/kg- day groups on LD 22. The samples were collected 24 hours af ter dosing. In the controls, serum PFOA was below the limits of quantitation (0.00528 μg/ml). Levels of PFOA found in female sera increased between the two dose groups; treated females had an average concentr ation of 0.37 + 0.0805 and 1.02 + 0.425 μg/ml, respectively for the 10 and 30 mg/kg-day dose groups.

3.8.2 F1 Generation No effects were reported at any dose level for the viability and lactation indices. No differences between treated and control groups were noted for the numbers of pups surviving per litter, the percentage of male pups, litter size and average pup body weight per litter at birth. At 30 mg/kg-day, one pup from one dam died prior to weaning on l actation day 1 (LD1). A dditionally, on lactation days 6 and 8, statistically signifi cant increases in the numbers of pups found dead were observed at 3 and 30 mg/kg-day. According to the study authors, th is was not considered to be treatment related because they did not occur in a dose-related manner and did not appear to affect any other measures of pup viability including numbers of surviving pups per litter and live litter size at weighing. An independent statistical analysis was conducted by US EPA (2002b). No significant differences were observed between dose groups and the respons e did not have any trend in dose. The authors did not present the mean pup body wei ghts for the male and female pups separately. Pup body weight on a per litter ba sis (sexes combined) was reduced throughout lactation in the 30 mg/kg-day group, and statis tical significance (p < 0.01) was achieved on days 1 (9.5%), 5 (9.6%), and 8 (10.5%). Of the pups necrop sied at weaning, no statistically significant, treatment-related differences were observed for the weights of the brai n, spleen and thymus and the ratios of these organ weights to the terminal body we ight and brain weight.

3.8.2.1 F1 Males Significant increases in treatment -related deaths (5/60 animals) were reported in F1 males in the high dose group of 30 mg/kg-day between days 2- 4 postweaning. One rat was moribund sacrificed on day 39 postweaning and another was found dead on day 107 postweaning. Statistically significant increases in clinical signs of toxicity were also observed in F1 males during most of the entire postweaning pe riod. These signs included an in creased incidence of annular constriction of the tail at all doses, with statistical significance at the 1, 10, and 30 mg/kg-day; a significant increase at 10 and 30 mg/kg-day in the number of male rats th at were emaciated; and a significant increase in the incidence of urine-stained abdominal fur, decreased motor activity, and abdominal distention at 30 mg/kg-day. Statistically significant reductions in body weight gain were observed at 10 and 30 mg/kg-day during days 8-15, 22-29, 29-36, 43-50, and 50-57 postweani ng. Body weight gains were also significantly reduced in the 30 mg /kg-day group on days 1-8, 15-22, 36-43, 57-64, and 64-70 postweaning. Body weights were significantly reduced inthe 10 mg/kg-day group beginning on postweaning day 36, and in the 30 mg/kg-day gr oup beginning on postweaning day 8. In the 3 mg/kg-day group,

159OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

mean body weight gain was signifi cantly reduced on days 43-50 and 57-64 postweaning, and mean body weights were significantly reduced on days 106 and 113 postweaning. In the 1 mg/kg-day group, mean body weight gain was significantly re duced on days 15-22 and 43-50 postweaning, and mean body weights were signi ficantly reduced on days 50, 57, 64, 70, 99, 106 and 113 postweaning. For all groups, there was a sign ificant, dose-related reduction in mean body weight gain for the entire dosing period (days 1-113). Ab solute food consumption values were significan tly reduced at 10 and 30 mg/kg-day during the enti re precohabitation period (days 1-70 postweaning), while relative food consumption values were significantly increased. Statistically significant (p < 0.01) delays in sexual maturation (the average day of preputial separation) were observed in high-dose an imals versus concurrent controls (52.2 days of age versus 48.5 days of age, respectively). No apparent effects were observe d on any of the mating or fertilityparameters including fertility and pregnancy indices (number of pregnancies per number of rats that mated and rats in cohabitation, respectively), the number of days to insemina te, the number of rats that mated, and the number of rats with confirmed mating da tes during the first week. No statistically significant, treatmentrelated effects were observed on any of the sperm parameters (motility, concentration, or morphology). Necroscopic examination revealed statistically significant treatment-relate d effects at 3, 10, and 30 mg/kg-day ranging from tan areas in the lateral and median lobes of the liver to moderate to slight dilation of the pelvis of one or both kidneys. Statistically significant, dose-related decreases in terminal body weights were observed in the 1, 3, 10, and 30 mg/kg-day dose groups(6%, 6%, 11%, and 22% decrease from controls, respectively; p < 0.01 at 1 and 3 mg/kg-day, p < 0.05 at 10 and 30 mg/kg-day). The abso lute and relative weights of the liver were statistically significant ly increased in a ll treated groups (p < 0.01) and was accompanied by histopathological changes. The absolute wei ghts of the left and/or right kidneys were statistically significantly increased in the 1 and 3 mg/kg-day dose groups and statisticallysignificantly decreased in the 30 mg/kg-day dose group. Organ we ight-to-terminal body weight and brain weight ratios for the kidney were statisti cally significantly incr eased in all treated groups. The biological significance of the effects on ki dney weight is unknown since hi stopathology was not conducted on that organ. All other organ weight changes observed (thymus, spleen, left adrenal, brain, prostate, seminal vesicles, testes, and epididymis) are prob ably due to decrements in body weight and not a reflection of target organ toxicity since the absolute weights of th ese organs was significantly reduced while the relative weights were either normal or significantly increased. However, the biological significance of the weight changes observed in the adrenal is unclear since hi stopathological changes were also observed. Histopathologic examination of the reproductive organs was unremarkable; however, treatmentrelated microscopic changes were observed in the adrenal glands of high- dose animals (cytoplasmic hypertrophy and vacuolation of the cells of the adrenal cortex) and in the liver of animals treated with 3, 10, and 30 mg/kg-day (hepato cellular hypertrophy). No other treatment-related effects were reported.

3.8.2.2 F1 Females A statistically significant increase in treatment-related mortality (6/60 animals) was observed in F1 females on postweaning days 2-8 at the highest dose of 30 mg/kg-day. No adverse clin ical signs of treatment-related toxicity were reported for any dose level during any time of the study period.

160OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Statistically significant decreases in body weights were observed in high-dose animals on days 8, 15, 22, 29, 50, and 57 postweaning, during precohabitati on (recorded on the day cohabitation began, when F1 generation rats were 92-106 days of ag e), and during gestation and lactation. Body weight gain was significantly reduced during days 1-8 a nd 8-15 postweaning. Statis tically significant decreases in absolute food consumption were observed during days 1-8, 8- 15, and 15-22 postweaning, during precohabitation and during ge station and lactation in anim als treated with 30 mg/kg-day. Relative food consumption values were comparable across all treated groups. Statistically significant (p < 0.01) delays in sexual maturation (the average day of vaginal patency) were observed in high-dose animals versus concurre nt controls (36.6 days of age versus 34.9 days of age, respectively). Prior to mating, the study authors noted a statistically si gnificant increase in the average numbers of estrous stages per 21 days in high-dose animals (5.4 versus 4.7 in controls). For th is calculation, the number of independent occurrences of estrus in the 21 days of observation was determined. This type of calculation can be used as a screen for effects on the estrous cycle, but a more detailed analysis should then be conducted to determine whether there is truly an e ffect. 3M (2002) recently completed an analysis that showed there were no effects on the estrous cycl e; there were no differences in the number of females with > 3 days of estrus or with > 4 days of diestrus in the control and high dose groups. Analyses co nducted by the US EPA (2002a) al so demonstrated that there were no differences in the estrous cycle among th e control and high dose groups. The cycles were evaluated as having either regular 4-5 day cycles, uneven cycling (defined as brief periods with irregular pattern) or periods of pr olonged diestrus (defined as 4-6 day diestrus periods) extended estrus (defined as 3 or 4 days of cornified smear s), possibly pseudopregnant, (defined as 6-greater days of leukocytes) or persistent estrus (defined as 5-or greater days of cornified smears). The two groups were not different in any of the parameters measured. Thus, the increase in the number of estrous stages per 21 days that was noted by the study authors is due to the way in which the calculation was done, and is not biologically meaningful. No effects on any of the mating a nd fertility parameters (numbers of days in cohabitation, numbers of rats that mated, fertility index, rats with confirmed mating dates during the first week of cohabitation and number of rats pregna nt per rats in cohabitation). All natural delivery observations were unaffected by treatment at any dose level. Numbers of dams delivering litters, the duration of gestation, averages for implantati on sites per delivered litter, the gestation index (number of dams with one or more liveborn pups/number of pregnant rats), the numbers of dams with stillborn pups, dams with all pups dying and liveborn and stillborn pups were comparable among treated and control groups. The terminal body weights and absolute and relati ve pituitary weights are shown in Table 47 No treatment-related effects were obs erved in the terminal body weights of the F1 females. The absolute weight of the pituitary, th e pituitary weight-to- terminal body weight ratio, and the pituitary weight-to-brain weight ratio were statistically significantly decreased at 3 mg/kg-day and higher. Since there is not a clear dose-re sponse relationship and histologic examination did not reveal any lesions, the biological significance of the pituitary weight data is problematic. No other differences were reported for the absolute weights or ratios for other organs evaluate d. No treatment-related effects were reported following macrosc opic and histopathologic examinations.

161OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 47 Summary of Body and Pi tuitary Weights of F1 Females

Group Terminal Body Weight (g) a Pituitary Weight (g) a Pituitary/ Body Weight (%) a Pituitary/Brain Weight (%) a

Control 322.9 + 23.4 0.017 + 0.004 5.46 + 1.6 0.84 + 0.21 1 mg/kg-day 321.7 + 24.2 0.016 + 0.003 5.0 + 0.86 0.76 + 0.14 3 mg/kg-day 329.2 + 21.5 0.015 + 0.003* 4.65 + 0.86* 0.74 + 0.14* 10 mg/kg-day 325.1 + 23.5 0.015 + 0.002* 4.72 + 0.90* 0.72 + 0.12* 30 mg/kg-day 315.7 + 20.9 0.015 + 0.003* 4.75 + 1.03* 0.72 + 0.16**

a- Mean +SD * Significantly different from control at p <0.05 ** Significantly different from control at p <0.01

3.8.3 F2 Generation No treatment-related adverse clinical signs were obs erved at any dose level. Dead or stillborn pups were noted in both the control and treated groups. The deaths occurred on lactation days 1-18 with the majority occurring on days 1-6. However, th ere was no dose-response relationship and therefore were unlikely related to treatment. Statistically significant increases (p < 0.01) in the number of pups found dead were observed on lactation day 1 in th e 3 and 10 mg/kg-day gr oups. According to the study authors, this was not considered to be treatment related because they did not occur in a doserelated manner and did not appear to affect any other measures of pup viability including numbers of surviving pups per litter and live litter size at weighing. An independent st atistical analysis was conducted by US EPA (2002b). No significant differe nces were observed between dose groups and the response did not have any trend in dose. No effects were reported at any dose level for the viability and lactation indices. No differences between treated and control groups were noted for the numbers of pups surviving per litter, the percentage of male pups, litter size and average pup body weight per litter when measured on LDs 1, 5, 8, 15, or 22. Anogenital distances measured for F2 male and female pups on LDs 1 and 22 were also comparable among the five dosage gr oups and did not differ significantly. Terminal body weights in F2 pups were not significan tly different from controls. Absolute weights of the brain, spleen and thymus and the ratios of these organ weights-to -terminal body weight and to brain weight were also comparable among tr eated and control groups. There were no treatmentrelated effects following necroscopic examination, with the exception of no milk in the stomach of the pups that were found dead.

3.8.4 Conclusions of the Reproductive Toxicity Study Dosing with APFO at 30 mg/kg-day resulted in a delay in the onset of sexual maturation in both male and female F1 offspring. The authors of the st udy contend that the delays in sexual maturation (preputial separation or vaginal pate ncy) observed in high-dose animals are due to the fact that these animals have a decreased gestationa l age, a variable which they have defined as the time in days from evidence of mating in the F0 generation until evidence of sexual maturation in the F1 generation. The authors state that gestat ional age appeared to be decrea sed in high-dose animals at the time of acquisition (the time when sexual maturati on was reached), which they believe meant the animals in that group were younger and more immature than the control group, in which there was no significant difference in sexual maturation.

162OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

In order to test this hypothesis, the authors covaried separately th e decreases in body weight and in gestational age with the delays in sexual maturation in order to determine whether or not body weights and gestational age were a contributing factor. When th e body weight was covaried with the time to sexual maturation, the time to sexual maturation showed a dose related delay that was statistically significant at p < 0.05. This suggests that the delay insexual maturation was partly related to body weight, but not entirely. When gest ational age was covaried with the time to sexual maturation, there was no significant difference in the time of onset of sexual maturation between controls and high-dose animals. This indicates th at the effect of delayed sexual maturation could possibly be attributed to decreased gestational age. While it is known and commonly accepted that change s in the body weights of offspring can affect the time to sexual maturation, whether or not gestati onal age, as defined by the authors, also affects the time of sexual maturation is purely speculative , especially since there were no data provided by the authors to support this relati onship. Additionally, covaring gesta tional age with time to sexual maturation is problematic from a statistical standpoi nt. Since there was no si gnificant change in the length of gestation at 30 mg/kg-day, based on the au thors’ definition of ‘gestational age’, the decreases in gestational age would have to be due mostly to changes in time to sexual maturation. Therefore, sexual maturation is essentially being covaried with itself. Still, even if a relationship between gestational age and time to sexual maturati on were shown, it merely offers an explanation for the observed delays in sexual maturation in high- dose animals, but does not diminish its significance. A variety of endpoints are evaluate d throughout different lifestages in a two generation reproductive toxicity study. Therefore, some of the endpoints may be indicativ e of developmen tal/reproductive toxicity, while others may be indicative of adult toxicity. The selection of developmental endpoints (and the appropriate L OAELs and NOAELs) for this study wa s based on the Agency’s Developmental Toxicity Risk Assessment Guidelines (E PA, 1991) and the Agency’s Reproductive Toxicity Risk Assessment Guidelines (EPA, 1996). According to the guidelines, the period of exposure for developmental toxicity is prior to conception in either parent, through pr enatal development and continuing until sexual maturation. In contrast, the period during which a developmental effect may be manifested includes the entire lifespan of th e organism. For selection of the developmental endpoints from the two generation re productive toxicity study, attent ion was focused on effects that were noted during the period of developmental e xposure. Thus, only effects that occurred up to sexual maturation were considered relevant for a ssessing developmental toxi city. Effects occurring after sexual maturation were considered relevant fo r assessing adult toxicitysince it was not possible to determine whether the effects were due to developmental and/ or adult exposures. Therefore, under the conditions of the study, the L OAEL for F0 parental males is 1 mg/kg-day, the lowest dose tested, based on significant increases in absolute and relative liver weight. A NOAEL for the F0 parental males could not be determined si nce the increases in liver weight were seen at all doses tested. The NOAEL for F0 parental fema les is 30 mg/kg-day, the highest dose tested. A variety of developmental/reprodu ctive effects were noted in th e F1 generation. During lactation, there was a significant reduction in F1 mean body weight on a litter basis (sexes combined) in the 30 mg/kg-day group. F1 males in the 10 and 30 mg /kg-day groups exhibited a significant reduction in body weight gain during days 8-50 postweaning, and body weights were significantly reduced in the 10 mg/kg-day group beginning on postweaning day 36, and in the 30 mg/kg-day beginning on postweaning day 8. F1 females in the 30 mg/kg-day group exhibited a significant reduction in body weight gain on days 1-15 postweaning, and in body weights beginning on day 8 postweaning. There was a significant increase in mortality mainly during the first few days after weaning, and a significant delay in the timing of sexual maturation for F1 males and females in the 30 mg/kg-day group. For F1 males, the LOAEL for developmental/rep roductive toxicity wa s 10 mg/kg-day, and the

163OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

NOAEL was 3 mg/kg-day. For F1 females, the L OAEL for developmental/reproductive toxicity was considered to be 30 mg/kg-da y, and the NOAEL was 10 mg/kg-day. The LOAEL for adult systemic toxicity in the F1 males is 1 mg/kg-day based on significant, doserelated decreases in body weights and body weight ga ins (observed prior to and during cohabitation and during the entire dosing period) , and in terminal body weights; and significant changes in absolute and relative liver weights. A NOAEL for the F1 males could not be dete rmined since these effects were seen at all doses tested. The NOAEL and LOAEL for adult systemic toxicityin the F1 females are 10 and 30 mg/kg-day, respectively, based on sta tistically significant decr eases in body weight a nd body weight gains. The NOAEL for developmental/reproductive toxici ty in the F2 offspring was 30 mg/kg-day. No treatment-related effects were observed at any dos es tested in the study. However, it should be noted that the F2 pups were sacrificed at weaning, and thus it was not possi ble to ascertain the potential post-weaning eff ects that were noted in the F1 generation.

Testicular and ovarian effects we re noted in the 2-year rat study (Sibinski, 1987; se e section 3.6.1).

3.9 Immunotoxicity Studies in Animals Four immunotoxicity studies of PFOA have been conducted in mice. In the first study, Yang et al. (2000) administered 0.02 % PFOA to male C57Bl/6 mi ce in the diet for 2, 5, 7, or 10 days. At the end of the feeding period, mice we re sacrificed and the liver, spleen, and thymus were weighed. The effect of PFOA administration on the cellularity , cell surface phenotype, a nd cell cycle of thymocytes and splenocytes was determined. In addition, effects of exposure of thymocytes and splenocytes to PFOA in vitro were examined. Administ ration of 0.02% PFOA for 2, 5, 7, or 10 days resulted in a significant increase, relative to controls , in liver weight, even at the earliest time point. Also, a decrease in body weight was observed. Follo wing five days of admi nistration, significant decreases in thymus and spleen weight were noted. After admini stration of 0.02% PFOA for 7 days, significant decreases (85% and 80%, respectively) in the total number of thymocytes and splenocytes were observed. In addition, the number of thymocytes expressing both CD4 and CD8 decreased by 95%; the number expressing neither CD4 and CD8 decreased by 57%; and the number expressing either CD4 or CD8 decreased by 64% and 72%, respectively. Fo r the splenocytes, both T cells (CD3) and B cells (CD19) decreased by 75% and 86% , respectively. Also, significant decreases in both CD4 helper and CD8 cytotoxic sp lenic T cells were observed. Upon administration of 0.02% PFOA to mice for 7 days, thymocyte prolifer ation was also inhibited, as detected by cell cycle flow cytometry analyses. In vitro studies s howed that there was spontaneous apoptosis occurring in splenocytes and thymocytes after 8 or 24 hours of culturin g in the presence of varying concentrations (50, 100, or 200 M) of PFOA. However, PFOA did not significantlyalter the cell cycle under these conditions. In order to examine the dose dependency of the effects, Yang et al. (2001) administered C57Bl/6 mice diets consisting of 0.001%-0.05% PFOA (w/w) for 10 days. For examining the time-course, a diet containing 0.02% PFOA was given for 2, 5, 7 or 10 days. Effects of withdrawal of PFOA were also studied. The results showed that, at higher doses, a significant decrease, relative to controls, in body weight was observed, although no other apparent si gns of toxicity such as sores, lethargy, and poor grooming were noticed. However, a significan t decrease in total wa ter intake was observed. Mice receiving dietary PFOA for 10 days experienced significant incr eases in liver weight and peroxisome proliferation, as measur ed by induction of acyl-CoA oxidase with lauroyl-CoA or palmitoyl-CoA as substrate. These increases started atthe lowest dose and reached their maximal values at a dose of 0.003-0.01%. In contrast , the weight decreases of the spleen and thymus began at a

164OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

higher dose (0.01%) with no maximum reached withthe doses given. The time course studies showed that increased liver weight s and peroxisome proliferation we re evident at the earliest time point examined. In contrast, si gnificant thymus and spleen we ight decreases required PFOA administration for a period of at least 5 days, foll owing which the spleen weight remained constant while the thymus weight continue d to decrease. However, upon pr olonged treatment for one month, no further decreases in thymus and spleen weight s were observed. In another set of experiments, animals received 0.02% PFOA for 7 days, and then they received normal chow for a period of 10 days. These recovery experiments showed that the animals rapidly recovered the body weight the second day after withdrawal of PFOA. However, the liver weight did no t return to normal even after 10 days of recovery. Thymus recovery started on day 2 and was completed by day 10. The spleen weights returned to normal by day 2 post-withdrawal . In addition, the changes in thymus and spleen weight upon PFOA treatment and withdrawal paralle led the changes in tota l thymocyte and splenocyte counts. Furthermore, flow cytometry cell cycle experiments s howed that the decrease in thymocyte number caused by PFOA treatment is due ma inly to inhibition of thymocyte proliferation. In contrast, PFOA treatment caused no cha nges in the cell cycle of splenocytes. A third feeding study (Yang et al. 2002a) was desi gned to examine the possi ble involvement of the peroxisome proliferator-activated receptor alpha (PPAR

 ) in the immunomodulation exerted by PFOA. This study made us e of transgenic PPAR

 null mice, which are homozygous with regards to a functional mutation in the PPAR

 gene. These mice do not exhibit pe roxisome proliferation or hepatomegaly and hepatocarcinogenesis even after e xposure to peroxisome proliferators. These mice were fed a diet consisting of 0.02% PFOA (w/w) fo r 7 days. At the end of the feeding period, mice were sacrificed and the liver, sp leen, and thymus were removed and weighed. The effect of PFOA on peroxisome proliferation, cell cycle, and lymphoproliferation was ascertained. The results showed that, in contra st to wild-type mice, feeding PPAR

 null mice PFOA resulted in no significant decrease in body wei ght. As expected, peroxisome proliferation, as measured by fatty acid oxidation, was totally lacking in PPAR

 null mice. Also in contrast to wild type mice, feeding PPAR α null mice PFOA resulted in no significant decrease in the weight of the spleen or the number of splenocytes. At the same time, the d ecrease in weight and cellularity of the thymus was attenuated, but not totally eliminated in the PPAR α null mice. In addition, the decreases in the size of the CD4+CD8+ population of thymus cells and the number of thymus cells in the S and G2/M phases of the cell cycle, which reflects inhib ition of proliferation, obse rved in wild type mice administered PFOA were much less extensive in PPAR α null mice. Finally, in contrast to wild type mice, PFOA treatment caused no significant change insplenocyte proliferation in response to mitogens in PPAR α null mice. A fourth feeding study (Yang etal. 2002b) was designed to exam ine the effects of PFOA on specific humoral immune responses in mice. For th is study, 0.02 % PFOA was administered to male C57Bl/6 mice for 10 days. Then the animals were examined, via plaque forming cell (PFC) and serum antibody assays, for their ability to generate an immune response to horse red blood cells (HRBCs). Ex vivo and in vitro splenic lymphocyte proliferation assays were also performed. The results showed that mice fed normal chow responde d to challenge with HRBCs with a strong humoral response, as measured by the PFC assay. In contrast, mice fed with PFOA responded to HRBC immunization with no increase in HRBC-speci fic PFCs, relative to unimmunized controls.However, in experiments where PFOA-treated mice received normal chow following HRBC immunization, there was a significant recovery of the numbers of specific PFCs stimulated. The suppression of the humoral immune response by PFOA was confirmed by analysis of the serum anti-HRBC response. In ex vivo experiments, splenocytes is olated from control mice responded to both ConA and LPS with lymphocyte proliferation, as measur ed by thymidine incorporation. However, treating mice with PFOA (0.02% for 7 days) attenuated the prol iferation. In a set of in vitro experiments,

165OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

PFOA (1- 200 M) added to the culture medium of splenocytes cultured from untreated mice did not cause an alteration of lymphocyte prolif eration in response to LPS or ConA.
Conclusions

In these studies there is evid ence that PFOA may be 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. 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 PPAR α was involved in causing the adverse effects to the immune system. Additional studies using conventional protoc ols are underway to confirm the results of these studies.

4 HAZARDS TO THE ENVIRONMENT

4.1 Aquatic Effects Acute Toxicity Test Results Concerning the acute toxicity of PFOA for aquatic organisms, experi mental data are available for three trophic levels (cf. Table 48). The tests were performed with ammonium perfluorooctanoate (APFO), the ammonium salt of PFOA.
A 96 h-LC 50 of 707 mg/l for Oncorhynchus mykiss was evaluated in a static test according to OECD TG 203 (CIT, 2003b). No analytical monitoring of th e test item was carried out as the substance was expected to be stable in water under the test conditions. Animals were exposed to nominal test concentrations between 31.3 and 10 00 mg/l. Transient sub-lethal e ffects (color change and abnormal respiratory rates) were observed in animals lying at the bottom of the aquarium at 250 and 500 mg/l. For both concentrations, th ese effects were only observed fr om 24 to 72 hours but not subsequently. An acute toxicity test with Daphnia magna was performed according to OECD TG 202 (CIT, 2003a). The daphnids were exposed to nominal APFO concentrations between 100 and 1000 mg/l in a static test system. No analyt ical monitoring of the te st item was carried out as the substance was expected to be stable in wate r under the test conditions. A 48 h-EC 50 of 480 mg/l for immobilization was calculated. The toxicity of APFO was ev aluated with the green algae Pseudokirchneriella subcapitata in a static test system according to OECD TG 201 (CIT , 2004a). The stability of the test substance was experimentally determined by ion chromatogra phy with electrochemical detection (conductimetry) at the beginning and at the end of the test. Recove ry rates were within ±20%, therefore the results are expressed as nominal values. Nominal APFO concentrations were in the range of 5.25 to 400 mg/l. The 72h and 96h-EC 50 for both biomass and growth rate were > 400 mg/l. NOECs for both biomass and growth rates were 200 mg /l after 72h, and 12.5 mg/l after 96h. Additional but less reliable tests with fish (96 h-LC 50 values from 300 to >1000 mg/l), daphnids (48 h-EC 50 values from 240 to 584 mg/l) and algae (72 h E r C 50 = 3330 mg/l) are available. The results are presented in Table 48.

166OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Chronic Toxicity Test Results Concerning prolonged and chronic t oxicity of PFOA towards aquatic species, experimental data are available for two trophic levels (cf. Table 48) . The tests were conducted with ammonium perfluorooctanoate (APFO), the ammonium salt of PFOA. For both tests the stability of the test substance in the medium was experimentally determ ined by ion chromatography with electrochemical detection (conductimetry). Analytical recovery rates were less than80% so the results were expressed as effective values. The prolonged toxicity of APFO to fish ( Oncorhynchus mykiss ) was tested under flow-through conditions in an ELS test accordi ng to OECD TG 210 (CIT, 2004b). The test started immediately after fertilization and was terminated 60 days after ha tching in the control at day 85. For the endpoint mortality of embryos (post hatch), larvae, and juveniles a NOEC of 40.0 mg/l (the highest concentration tested) was obtained. Also a NOEC of 40 mg/l for sub-lethal endpoints as mean weight and mean length was determined. A 21-d-reproduction study with Daphnia magna according to OECD TG 211 was conducted with APFO in a semi static test system (CIT, 2003c). Test solutions were renewed three times a week. A 21 d-NOEC of 20 mg/l and a LOEC of 44 mg/l for reproduction cap acity (number of living neonates per parent alive) was obtained. The 21 d-NOEC based on length of parent animals was 44 mg/l and based on survival of parent animals was 88.6 mg/l.

Table 48 Toxicity of APFO to aquatic species

Species Test type Parameter Effects Reference IUCLID Oncorhynchus mykiss Static 96 h-LC 50 = 707 mg/l (n) CIT, 2003b 4.1
Oncorhynchus mykiss Static 96 h-LC 50 = 800 mg/l (e) DuPont, 1999 4.1

Lepomis macrochirus Static 96 h-LC 50 = 634 mg/l (n) DuPont, 1994 4.1

Pimephales promelas Static 96 h-LC 50 140 mg/l (n) * 3M, 1996d 4.1

Pimephales promelas Static 96 h-LC 50 > 301 mg/l (n) 3M, 1987 4.1

Pimephales promelas Static 96 h-LC 50 = 766 mg/l (n) 3M, 1980 4.1

Pimephales promelas Static 96 h-LC 50 > 1000 mg/l (n) 3M, 1990c 4.1

Pimephales promelas Static 96 h-LC 50 2470 mg/l (n) 3M, 1996a 4.1

Oncorhynchus mykiss (early life stage) Flowthrough Mortality of embryos, larvae, juveniles 85 d-NOEC 85 d-LOEC = 40 mg/l (e) > 40 mg/l (e) CIT, 2004b 4.5.1

Pimephales promelas (eggs, larvae) Flowthrough Hatching success, 30 d post-hatch success of larvae 30 d-NOEC 30 d-LOEC = 100 mg/l (n) > 100 mg/l (n) EG & G, Bionomics, 1978 4.5.1

167OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

168 Species Test type Parameter Effects Reference IUCLID Daphnia magna Static Immobilisation 48 h-EC 50 = 480 mg/l (n) CIT, 2003a 4.2

Daphnia magna Static 48 h-EC 50 = 584 mg/l (n) 3M, 1990b 4.2 Daphnia magna Static 48 h-EC 50 = 240 mg/l (n) 3M, 1996b 4.2 Daphnia magna Static 48 h-EC 50 = 360 mg/l (n) 3M, 1996c 4.2

Daphnia magna Semistatic Reproduction 21 d-NOEC 21 d-LOEC = 20 mg/l (e) = 44 mg/l (e) CIT, 2003c 4.5.2

Pseudokirchneriella subcapitata Static Biomass, growth rate 72 h-EC 50 72 h-NOEC
96 h-EC 50 96 h-NOEC > 400 mg/l (n) = 200 mg/l (n) > 400 mg/l (n) = 12.5 mg/l (n) CIT, 2004a 4.3

Selenastrum capricornutum Static 72 h-E r C 50 =3330 mg/l (n) 3M, 1996e 4.3

(e): effective concentration (n): nominal concentration * : low pH values in high test concentrations may have had an adverse effect on survival Toxicity to Microorganisms The toxicity of APFO to activated sludge was evaluated in seve ral respiratory inhibition tests. The results of the toxicity tests with microorganisms are listed in Table 49.

Table 49 Toxicity of APFO to microorganisms
Species Parameter Effects Reference NOEC (7 min) = 1000 1* 3M, 1980a; US EPA, 2002
EC 50 (30 min) > 1000 1

EC 50 (30 min) > 500 2

EC 50 (30 min) > 664 3 US EPA, 2002

EC 50 (3 h) > 1000 3M, 1987; 3M, 1990 activated sludge

EC 50 (3 h) > 3320 3M, 1996 1 Test substance 100% APFO * Test documentation insufficient for assessment, report lacks record of methodology and analytical measurement of test substance concentrations in the test solutions. 2 Test substance APFO in 50% water, va lue adjusted to 100% active ingredient 3 Test substance APFO in 80% water, va lue adjusted to 100% active ingredient

4.2 Terrestrial Effects Acute Toxicity Test Results Tominaga et al. (2004) conducted test s with the soil dwelling nematode Caenorhabditis elegans . 20 worms each were exposed to different nominal concentrations (0.01 to 5.0 mM [4.14 to 2070 mg/l)OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

of PFOA on NGM agar plates. The test solutions were poured on plates before the worms were added. After 1, 2, 3, 4, 24 and 48 hours the total nu mber of worms and the number of active worms on each plate was counted. These experiments were ca rried out in triplicate and repeated at least 3 times. A 24 h-EC 50 of 2.75 mM (1138 mg/l) and a 48 h-EC 50 on activity of the worm of 2.35 mM (973 mg/l) was calculated usi ng the PROBIT analysis. Chronic Toxicity Test Results In a multi-generation test the effects of PFOA on the fecundity of the soil dwelling nematode Caenorhabditis elegans was investigated (Tominaga et al ., 2004). The animals were exposed to concentrations of 100 pM, 1 nM, 10 nM and 100 μM [equivalent to 4.1x10 -5 , 4.1x10 -4 , 4.1x10 -3

, 41.4 mg/l, respectively]. L1-larvae of the first ge neration were added to NGM agar plates. When worms grew to L4 larvae, one worm was picked up and transferred to a new plate of the same composition and incubated at 16°C. The number of worms and eggs were counted in 24 h-intervals. The second- generation worms were allo wed to grow to L4 larvae on the same plate, and then again one worm was picked up and transferred to a new plat e of the same composition. These steps were repeated until the fifth generation was cultured. These experiments were carried out in triplicate and repeated at least 3 times. At the highest concentration tested (100 μM = 41.4 mg/l) the worm abundances from first to second generation showed an ‘extreme’ decrease, so sub-cu lturing was discontinued at this exposure level. In all other concentrations wo rm abundances and egg numbers dur ing the first generation did not differ significantly from controls. In the forth generation a signifi cant decrease in worm abundances and egg production was observed at 10 nM (= 4.1x10 -3 mg/l). For the endpoints worm abundance and egg production a NOEC of 1 nM (4.1x10 -4 mg/l) and a LOEC of 10 nM (4.1x10 -3 mg/l) was maintained. In the acute and multi-generation study with Caenorhabditis elegans the culture conditions and the exposure to PFOA on NGM agar plates ar e not relevant for the environment.
Conclusion The acute toxicity of PFOA to Caenorhabditis elegans is low. After exposur e for 4 generations the most sensitive sublethal endpoint s were abundance and egg productio ns with NOEC values about 2.3x10 -6 lower than acute the EC 50 -value.

4.3 Other Environmental Effects Toxicity to Sediment Dwelling Organisms In a range-finding test with PFOA w ith the sediment dwelling organism Chironomus tentans a 10 dEC 50 and LC 50 > 100 mg/l (the highest concentration test ed) was obtained (McDonald et al., 2004). No significant impacts on survival and growth were observed. The test was conducted according to US EPA-600-R99-064 and ASTM E-729-96 guidelines . The animals were e xposed via overlaying water. Test solutions were renewed every 48 h beginning on day 2. No an alytical monitoring was conducted in this range finding test.
Microcosm Studies The impact of sodium perfluor ooctanic acid (CAS No. 335-95-5) on the structure of the zooplankton community was evaluated in indoor microcosms over a period of 35 days (Sanderson et al., 2002, 2003, 2004). PVC aquariums with a volume of 30 l were randomly filled with sediment, animals, phytoplankton (ca. 10 3 cells / ml) and clean water from the field. Th e zooplankton community consisted of the following representative species: Cyclops diaptomus , Cyclops strenuous , Cyclops

169OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

canthocampus staphylinus , Daphnia magna , Keratella quadrata , Phyllopoda sp., Echninorhynchus sp., Ostracoda sp., and Rotifera sp. Also occasional macrophytes ( Elodea canadensis , Myriophyllum spicatum ) and larger invertebrates ( Ephemeroptera sp., Asselus aquaticus ) were present. Single application of 1, 10, 20, 30, 70 mg/l (nominal; based on the anion) in five replicates for each concentration and controls were tested. Only temporary statistically significant structural fluctuations were observe d. The fluctuations were always followed by rapid recovery and therefore, there are no ecologi cal significant effects in this study. Daphnia magna seemed to be the most sensitive specie s of the pelagic compartment with a LOEC of 20 mg/l (NOEC = 10 mg/l) 24 h post-treatment. In contrast, the smaller and more robust species of Rotifera sp. increased in abundance throughout the study at the two highest treatment concentrations of 30 mg/l and 70 mg/l. Total zooplankton abundance showed no signifi cant trends. Zooplankt on abundance trends observed in the 70 mg/l-microcosms were consistently lower than those found in all other concentrations throughout the study. Total zooplankton abundance was significantly (p<0,01) increased at 1, 10, 20 and 30 mg/l, but not at 70 mg/l. Species richness was highest in low treatment mi crocosms and lowest in high treatment microcosms. The overall species ric hness was significantly (p<0,05) reduced at 10, 20, 30 and 70 mg/l, indicating a simplification of community structure. A shift fr om a more diverse community with more total zooplankton species to wards less richness dominated by smaller zooplankton species like Rotifera sp. could be observed.

A NOEC communitywas not determined because of high within -treatment variability and low statistical power due to high inter-replicate variance and relatively small effect sizes. A LOEC species richness of 10 mg/l can be deduced. The authors of the study presented a tentativ e order of descriptor sensitivity as follows: Daphnia magna > richness > Cyclops canthocampus staphylinus > Cyclops diaptomus > total zooplankton > Rotifera sp. Oakes et al. (2004) investigated the reproductive impairment and biochemical changes in fathead minnow ( Pimephales promelas ) exposed for 39 days to varying c oncentrations of perfluorooctanoic acid (PFOA) under outdoor microcosm conditions. Test concentrations used were 0.3, 1.0, 30 and 100 mg/l (0.27, 0.65, 23.9 and 74.1 mg/l as TWA [time-we ighted averages] of measured conentrations). No mortality was associated with PFOA exposure. Si gnificant declines in circulating plasma steroids (testosterone) were observed in both sexes, but these were accompanied by only limited increases in time to first oviposit ion and decreases in overall egg produc tion in females. Only modest changes were observed in condition factor (females) and in relative liver size (in males); gonad sizes were unchanged in both sexes. Peroxisome proliferation, as quantified by fatty acyl-CoA oxidase (FAO) activity, was elevated with low PFOA concentrations, but attenuated with exposure to higher PFOA doses. Little evidence was seen for differential induction of peroxisome- associatedenzyme activity with sex. Oxidative stress, as quantified by the 2-thiobarbituric acid reactive substances (TBARS) assay, was only modestly influe nced by PFOA exposure and is not a significant consequence of FAO activity in the fathead minnow. The NOEC for survival, was > 100 mg/l (nominal); the NOEC for time to first oviposition was 24 mg/l (effective); the NOEC for male plasma 11-keto testosterone and testos terone was 0.27 mg/l (effective).

170OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

The toxicity of PFOA to the two macrophyte species Myriophyllum spicatum and M. sibiricum was evaluated in an outdoor microcosm study (Hanson et al., 2005) . Replicate microcosms (n=3) were treated with 0.3, 1.0, 30 and 100 mg/l PFOA as sodium salt (0.27, 0.65, 23.9 and 74.1 mg/l as TWA (time-weighted averages) of measured concentrat ions) and assessed at regular intervals over 35 days. At each sampling interval (at day 7, 14, 21 and 35) two plants of eac h species were randomly removed from the microcosms and examined for several endpoints. These were growth (plant length), biomass (wet mass/dry mass) , root number (primary roots from the plant stem), primary root length (total and longest), number of nodes. Pigment concentrations (chlorophyll a/b and carotenoid concentrations) were also examined. The two species were similar in their sensitivityto PFOA under the simula ted field conditions. No statistically significant differences or concentration-response trends were noted after 7 days of exposure to PFOA in both species. A 35d-NOEC of 23.9 mg/l and a 35d-LOEC of 74.1 mg/l was determined. For M. spicatum a 35d-EC 10 of 5.7 mg/l and for M. sibiricum a 35d-EC 10 of 8.4 mg/l was calculated. Effect values are based on TWA of measured concentrations.

171OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

6 REFERENCES 3M (1977). Ready Biodegradation of FC-143 (BOD/COD/TOC). Lab Re quest Number 3844 3M Company Environmental Laboratory. St. Paul, MN U.S. Environm ental Protection Agency Administrative Record 226-0487 (July 14). 3M (1978) Reiner EA, Fate of Fluorochemicals in the Environment. Biodegradation Studies of Fluorocarbons III, Project number 9970612613. Envir onmental Laboratory, St. Paul, MN. U.S. Environmental Protection Agency Admini strative Record 226-0489 (July 19). 3M (1978b) Unpublished Data, Welsh SK, 3M Co. Environmental Laboratory Technical Report Summary, Adsorption of FC-95 and FC-143 on Soil (February 27). Environmental Laboratory. 3M Company Project 9970612633: Fate of Fluorochemicals, Report Number1. St. Paul, MN. U.S. Environmental Protection Agency Administrative Record 226-0488 (February 27). 3M (1979) Todd JW FC-143 Photolysis Study Using Simulated Sunlight. Project 9776750202, 3M Company Environmental Laboratory Technical Report No. 002. February 2., FC-143 Photolysis Study Using Simulated Sunlight U.S. Environmental Protection Agency Administrative Record 226-0490. 3M (1980). Ready Biodegradation of FC-143(B OD/COD). Lab Request Nu mber 5625S. Environmental Laboratory. St. Paul, MN. U.S. Environm ental Protection Agency Administrative Record 226-0492 (April 11). 3M (1980a). Unpublished Data, 3M Co. Envir onmental Laboratory Lab Request number 5625S, Activated Sludge Respiration Inhibition (June 4), U.S. Environmental Protection Agency Administrative Record 226-0505. 3M (1980b). Environmental Laboratory Acute Toxici ty Testing: FC-14 (May 8). Unpublished Data. U.S. Environmental Protection Agency Administrative Record 226-0504. 3M (1985). Ready Biodegradation of FX- 1001 (BOD/COD). Lab Request No.C1006. Environmental Laboratory. St. Paul, MN. U.S. Environm ental Protection Agency Administrative Record 226-0494 (February 14). 3M (1987). Fluorad® Fluorochemical Surfactant FC-143. 3M Company Technical Bulletin. 3M Co. Environmental Laboratory (1987). Unpublished Data, 3M Co. Environmental Laboratory Lab Request number E1282, , Activated Sludge Resp iration Inhibition (April 20), U.S. Environmental Protection Agency Administrative Record 226-0510. 3M (1987). Unpublished Data, Microbi cs Microtox Test (April 29) U.S. Environmental Protection Agency Administrative Record 226-0511. 3M (1987b). Environmental Laboratory, 96-Hour Acute Static Toxicity to Fathead Minnow,Pimephales promelas (May 1). Unpublished Data, U.S. Environmental Protection Agency Administrative Record 226-0513. 3M (1990). Unpublished Data, 3M Co. Environmenta l Laboratory., Activated Sludge Respiration Inhibition Test (February 16), U. S. Environmental Protection Agency Administrative Record 226- 0514. 3M (1990b). Study No. 9013-3, Static Acute Toxicity to the Daphnid, Daphnia magna (April 9). EnviroSystems, Inc. Unpublished Data U.S. Envi ronmental Protection Agency Administrative Record 226-0517.

172OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

3M (1990c). Study No. 9014-3, Static Acut e Toxicity to the Fathead Minnow, Pimephales promelas (March 15). EnviroSystems Inc. U.S. Environm ental Protection Agency Administrative Record 226-0516. 3M (1993). Impinger Studies of Volatility of FC -95 and FC-143. 3M Laboratories, St. Paul, MN. 3M Lab Request Number L3306. U.S. Environmental Protection Agency Administrative Record 226-1030a-024. 3M (1995). Unpublished Data, 3M Co. Environmen tal Laboratory, , Assessment of Bioaccumulative Properties of Ammonium Perfluorooctanoic Acid: Static Fish Test (May 31) U.S. Environmental Protection Agency Administrative Record 226-0496. 3M (1996). Unpublished Data, 3M Co. Environmenta l Laboratory., Activated Sludge Respiration Inhibition Test (April 25) U.S. Environmental Pr otection Agency Administrative Record 226-0524. 3M (1996a). Study No. 1031-TH, Acute T oxicity to the Fathead Minnow, Pimephales promelas (May 15). T.R. Wilbury Laborator ies. Inc. U.S. Environmental Protection Agency Administrative Record 226-0525. 3M (1996b). Study No. 1030-TH, Acute Toxicity to the Daphnid, Daphnia magna (May 15). T.R. Wilbury Laboratories, Inc. Unpubl ished Data U.S. Environmental Protection Agency Administrative Record 226-0527.
3M (1996c) Unpublished Data, Study No. 892-TH, Acute Toxicity to the Daphnid, Daphnia magna (March 7). T.R. Wilbury Laboratories. Inc. Unpublished Data, U.S. EPA Administrative Record 226-0520. 3M (1996d) Acute Toxicity to the Fathead Minnow, Pimephales pr omelas (January 24). T. R. Wilbury Laboratories, Inc. Unpubl ished Data), U.S. Environmenta l Protection Agency Administrative Record 226-0519. 3M (1996e). Study No. 1029-TH, Grow th and Reproduction Toxicity Te st to the Freshwater Alga, Selenastrum capricornutum (May 15). T.R. Wilbury Laboratories, Inc. Unpublished Data, U.S. Environmental Protection Agency Administrative Record 226-0526. 3M (1999a). 3M Company Letter (dated May 26, 1999) re: TSCA 8(e) Supplemental Notice: Sulfonate-based and carboxylic -based fluorochemicals–Dock et Nos. 8EHQ-1180-373; 8EHQ-1180- 374; 8EHQ-0381-0394; 8EHQ-0699-373. 3M (1999b). The science of organic fluorochemistry and perfluorooctane sulfonate: Current summary of human sera, health, and toxicol ogy data. 3M Company February 5, 1999. 8EHQ-0299-373. 3M (1999c) Fluorochemical Use, Distribution and Release Overvi ew. Prepared by the 3M Company for USEPA OPPT, May 26, 1999. U.S. Environm ental Protection Agency Administrative Record 226-0550. 3M (2000a). “Voluntary Use and Exposure Informa tion Profile Perfluorooc tanic Acid and Salts” 3M Company. U.S. Environmental Protection Agen cy, Washington, D. C. U.S. Environmental Protection Agency Administrative Record 226-0595. 3M (2000b). Phase-Out Plan for POSF-Based Pr oducts. US EPA Administrative Record AR226- 0600 Publicly available at www.regulations.gov. 3M (2001a). Hydrolysis Reactions of Perfluor ooctanoic Acid (PFOA). 3M Lab Request Number E00-1851. U.S. Environmental Protection Agency Administrative Record 226-1030a090 (March 30).

173OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

3M (2001b). Hatfield T, Screening Studies on the Aqueous Photolytic Degradation of Perfluorooctanoic acid [PFOA] Lab request number E00-2192. Environmental Laboratory, St. Paul, MN. U.S. Environmental Protection Agency Administrative Record 22 6-1030a091 (April 20). 3M (2001c). Environmental Monito ring – Multi-City Study: Water, Sludge, Sediment, POTW Effluent and Landfill Leachate Samples. Executive Summary. U.S. Environmental Protection Agency Administrative Record 226-1030a111 (June 25). 3M (2003a). Environmental, Health and Safety Meas ures Relating to Perfluorooctanoic Acid and its Salts (PFOA). 3M Company letter. U.S. Envi ronmental Protection Agency Record OPPTS 2003- 0012-0007 (March 13). 3M (2003b) Submission of monitoring data pur suant to the 3M LOI dated March 13, 2003 and APFO users LOI dated March 14, 2003. U.S. Environmental Protect ion Agency Administrative Record 226-1482. 3M Company (2003c). Environmental, Health and Safety Measures Relating to Perfluorooctanoic Acid and its Salts (PFOA). U.S. Environmental Protection Agency Administrative Record 226- 1303. 3M (2003d) Inherent aerobic biodegradability of fluoroaliphatic polymeric ester of [confid. buisiness information], E02-0913. 3M (2005a). Analysis of PFBS, PFHS, PFOS, and PFOA in Water Samples Collected at 3M Guin. Amended Analytical Report E05-0662. U.S. Envir onmental Protection Agency Administrative Record 226-3571. 3M (2005b). Perfluorooctanoic Acid (PFOA) Site Related Environmental Assessment Program 3M and Dyneon LLC Decatur, Alab ama. EPA-HQ-OPPT-2004-0112-0042. Abdellatif AG, Preat V, Taper HS, and RoberfroidM (1991). The modulation of rat liver carcinogenesis by perfluorooctanoic ac id, a peroxisome proliferat or. Toxicol Appl Pharmacol 111 ,530-537. Alexander BH (2001a). Mortality study of workers employed at the 3M Cotta ge Grove facility. Final Report. April 26, 2001. Division of Environmental and Occupati onal Health, School of Public Health, University of Minnesota. U.S. Environm ental Protection Agency Administrative Record 226-1030a018. Alexander BH (2001b). Mortality st udy of workers employed at the 3M Decatur faci lity. Final Report. April 26, 2001. Division of Environmental and Occupational Health, School of Public Health, University of Minnesota. U.S. Environmental Protection Agency Administrative Record 226- 1030a019. Alzaga R and Bayona JM (2004). Determination of perfluorocarboxylic acids in aqueous matrices by ion-pair solid-phase microextraction–in-po rt derivatization–gas ch romatography–negative ion chemical ionization mass spectrometry. J Chromatogr A 1042 , 155-162. APME (2002). Association of Plastic Manufacturer s in Europe. The occurrence of perfluorooctanoate in environmental, dietary and human samp les. U.S. Environmental Protection Agency Administrative Record 226-1055, January 11. APME [Association of Plastic Manufacturer s in Europe] (2003). Dekleva LA, Adsorption/Desorption of Ammonium Pe rfluorooctanoate to Soil (OECD 106). DuPont Central Research & Development, Environmental & Microbiological Sciences & Engineering. Newark, DE (April 17).. DuPont EMSE Report Number EMSER 17-03.

174OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Armitage J, Cousins IT, Buck RC, Prevedouros K, Russell MH, MacLeod M and Korzeniowski SH (2006). Modeling Global Scale Fate and Transport of Perfluorooctanoate Emitted from Direct Sources. Environ Sci Tech web published 10/10/2006. Asahi Glass Co. (1975). Processing for producing perfluorocarboxylic aci d. British Patent 1473807. ATSDR (2005). 3M Chemolite Perfluorochemical Releases at the 3M Cottage Grove Facility City of Cottage Grove, Washington County, Minnesota EPA facility ID: MND006172969. Banks RE, Smart BE and Tatlow JC (1994). Organof luorine Chemistry: Principles and Commercial Applications. Plenum Press, New York. Barton DA, Butler LE, Zarzecki CJ, Flaherty JM, and Kaiser MA (2005). Characterizing perfluorooctanoate in ambient air near the fence lin e of a manufacturing faci lity: comparing modeled and monitored values. J Air Waste Managem Assoc, Accepted for publication pending revision. Barton DA, Butler LE, Zarzecki CJ, Flaherty JM, and Kaiser MA (2006). Characterizing perfluorooctanoate in ambient air near the fence line of a manufacturing facility: comparing modelled and monitored values. J Air Waste Manag Assoc. 56 (1), 48-55. Begley TH, White K, Honigfort P, Twaroski ML , Neches R, and Walker RA (2005). Perfluorochemicals: Potential sources of and migra tion from food packaging. Food Addit Contam 22 , 1023- 1031. Behar B and Stein G, (1966). Science 154, p.1012. Beilstein (2005). Handbook of Organic Chem istry (online). Request January 12. Berger U, Järnberg U, and Kalle nborn R (2004). Perfluorinated alky lated substances (PFAS) in the European Nordic environment. Organohalogen Compounds 66 , 4046-4052. Berger U and Haukas M (2005). Validation of a screening method based on liquid chromatography coupled with high-resolution mass spectrometry for an alysis of perfluoroalk ylated substances in biota. J Chromatography A 1081 , 210-217 Berthiaume J and Wallace KB ( 2002). Perfluorooctanoate, perfl uorooctanesulfonate, and N-ethylperfluorooctanesulfonamido ethanol; peroxisome proliferation and mitochondrial biogenesis. Toxicol Lett 129 , 23-32. Biegel LB, Liu RCM, Hurtt ME, and Cook JC (1995). Effects of a mmonium perfluorooctanate on Leydig cell function: in vitro, in vivo, and ex vivo studi es. Toxicol Appl Pharmacol 134 , 18-25. Biegel LB, Hurtt ME, Frame SR, O’Conner JC, and Cook JC (2001). Mechanisms of extrahepatic tumor induction by peroxisome proliferat ors in male CD rats. Toxicol Sci 60 , 44-55. Billot AR (2005) Perfluorochemical residential exposure data for Washington County, Minnesota. USAEPA Docket AR226-3563 Blume HP and Ahlsdorf B (1993). Prediction of pestic ide behavior in soil by means of simple field tests. Ecotoxicol Environ Saf 26 , 313 – 332 Bluist S, Cherrington NJ, Choudhuri J, Hartley DP, and Klaassen CD (2002). Gender-specific and developmental influences on the expression of rat organic anion transporters. J Pharmacol Exper Ther 301 , 145-151. Boese, W, Ho C, Knaup W, KochV (2005). Biodegradation potential of a Clariant fluoroteolomerbased acrylate polymer – Results from a test on inherent biodegradability(OECD 302B). Poster

175OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

presented at Fluoros: Internati onal Symposium on Fluorin ated Alkyl Organic in the Environment, August 18-20, 2005. Toronto, Canada. Boit H-G (1975). Beilstein Handbook of Organic Ch emistry. Springer-Verlag, Berlin. 4th Work, Volume 2, Part 2, page 994. Bossi R, Riget FF, Dietz R, Sonne C, Fauser P, Dam M, and Vorkamp K (2005). Preliminary screening of perfluorooctane sulf onate (PFOS) and other fluorochemicals in fish, birds and marine mammals from Greenland and the Faeroe Islands. Environ Pollut 136 , 323-329. Boulanger B, Vargo J, Schnoor JL , and Hornbuckle KC (2004). Dete ction of perfluorooctane surfactants in Great Lakes Wa ter. Environ Sci Technol 38 , 4064-4070. Boulanger B, Vargo JD, Schnoor JL, and Hornbuc kle KC (2005). Evaluation of perfluorooctane surfactants in a wastewater treatment system and in a commercial surface protection product. Environ Sci Technol 39 , 5524-5530. Boulanger, B., Peck, A.M., Schnoor, J.L., and Hornbuckle, K.C. (2005b). Mass Budget of Perfluorooctane Surfactants in Lake Onta rio. Environmental Science and Technology 39 : 74-79

. Boyd SA (1993a). Review of Tec hnical Report Summary: Adsorption of FC 95 and FC 143 in Soil. Michigan State University. U.S. Environmenta l Protection Agency Administrative Record 226- 0488 (May 19). Boyd SA (1993b). Review of Technical Noteb ook. Soil Thin Layer Chromatography. Michigan State University Number 48277, 30. U.S. Environmental Protection Agency Administrative Record 226-1030a089. Brace NO (1962). J Org Chem 27 , 4491. Burris JM, Olsen G, Simpson C, and Mandel J (2000). Determination of serum half-lives of several fluorochemicals. 3M Company. Interim Report #1. June 8, 2000. U.S. Environmental Protection Agency Administrative Record 226-0611. Burris JM, Lundberg JK, Olsen GW, Simpson D, and Mandel G (2002). Determination of serum half-lives of several fluoroch emicals. 3M Company. Interim Report #2. January 11, 2002. U.S. Environmental Protection Agency Administrative Record 226-1086. Butenhoff J, Costa G, Elcombe C, Farrar D, HansenK, Iwai H, Jung R, Kennedy G, Lieder P, Olsen G, and Thomford P (2002). Toxicity of a mmonium perfluorooctanoate in male cynomolgus monkeys after oral dosing for 6 months. Toxicol Sci 69 , 244-257. Butenhoff JL, Kennedy GL, Frame SR, O’Conner JC, and York RG (2004a). The reproductive toxicology of ammonium perfluorooctanoate (APFO) in the rat. Toxicology 196 , 95-116. Butenhoff JL, Kennedy GL Jr, Hinder liter PM, Lieder PH, Jung R, Hansen KJ, Gorman GS, Noker PE, and Thomford PJ (2004b). Pharmacokinetics of perfluorooctanoate in cynomolgus monkeys. Toxicol Sci 82 , 394-406.
Cabana, G., and Rasmussen, J.B. 1994. Modelling food chain structure and contaminant bioaccumulation using stable nitrogen isotopes. Nature (Lond.), 372: 255.257. Calafat AM, Kuklenyik Z, Reidy JA, and Needham LL (2006). Perfluorochemicals in United States residents in 2001 and 2002. Environ Sci Technol 40 2128-2134.

176OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Calafat AM, Needham LL, Kuklenyik Z, Reidy JA, Tully JS, Aguilar-Villalobos M, and Naeher LP (2005). Perfluorinated chemicals in selected resi dents of the American continent. Chemosphere (Epub Oct 5). Calfours J and Stilbs P (1985). Solubilization in sodium perfluorooctanoate micelles: a multicomponent self-diffusion study. Colloid Interface Sci 103 , 332-336. Caliebe C, Gerwinski W, Hühnerfuss H, and Theoba ld N (2004). Occurrence ofperfluorinated organic acids in the water of the North Sea. Organohalogen Compounds 66 , 4074-4078. Cameron RG, Imaida K, Tsuda H, and Ito N (1982). Promotive effects of steroids and bile acids on hepatocarcinogenesis initiated by diethylnitrosamine. Cancer Res 42 , 2426-2428. Cattley RC, DeLuca J, Elcombe C, Fenner-Crisp P, Lake BG, Marsman DS, Pastoor TA, Popp JA, Robinson DE, Schwetz B, Tugwood J, and Wahli W (1998). Do peroxisome proliferating compounds pose a hepatocarcinogenic risk to humans? Regul Toxicol Pharmacol 27 , 47-60. Chandra M, Riley MGI, and Johnson DE (1992). Spontaneous neoplasms in aged sprague-dawley rats. Arch Toxicol 66, 496-502. Chapin RE, Stevens JT, Hughes CL, Kelce WR, Hess RA, and Daston GP (1996). Symposium overview: endocrine modulation of reproduction. Fund Appl Toxicol 29 , 1-17. Chiappero MS, Malanca FE, Argue ello GA, Wooldridge ST, Hurl ey MD, Ball JC, Wallington TJ, Waterland RL, Buck RC (2006) Atmospheric Chemistry of Pe rfluoroaldehydes (CxF2x+1CHO) and Fluorotelomer Aldehydes (CxF2x+1CH2CHO): Quantification of the Important Role of Photolysis. J Phys Chem A, 110(43), 11944-11953. Cheung C, Akiyama TE, Ward JM, Nicol CJ, Feig enbaum L, Vinson C, and Gonzalez FJ (2004). Diminished hepatocellular proliferation in mice humanized for the nuclear receptor peroxisome proliferator-activated receptor α . Cancer Res 64 , 3849-3854. Christopher B and Marias AJ (1977). 28-Day oral toxicity study withFC-143 in albino mice, Final Report, Industrial Bio-Test Laboratories, Inc. Study No. 8532-10655, 3M Reference No. T1742CoC, Lot 269. CIT (2003a). Acute toxicity in Daphnia magna under static conditions . CIT/Study No. 22654 EAD/ Ammonium perfluorooctanoate (A PFO)/APME. Unpublished report. CIT (2003b). Acute toxicity in the rainbow trout under static conditions . CIT/Study No. 22655 EAP/ Ammonium perfluorooctanoate (APFO)/APME. Unpublished report. CIT (2003c). Daphnia magna reproduction test. CIT/Study N o. 22658 ECP/ Ammonium perfluorooctanoate (APFO)/APME . Unpublished report. CIT (2004a). Algal Inhibition Test. CIT/Study No. 23685 EAA/ Am monium perfluorooctanoate (APFO)/APME. Unpublished report. CIT (2004b). Early-life stage toxici ty in Rainbow trout under flow -through condition s. CIT/Study No. 22659 ECP/ Ammonium perfluorooctanoate (APFO)/APME. Unpublished report. Clegg ED, Cook JC, Chapin RE, Foster PMD, and Daston GP (1997). Le ydig cell hyperplasia and adenoma formation: mechanisms and re levance to humans. Reproduct Toxicol 11 , 107-121. Cook JC, Hurtt ME, Frame SR, and Biegel LB (1994) . Mechanisms of extrahepatic tumor induction by peroxisome proliferators in Crl:CD BR(CD) rats. Toxicologist 14 , 301 (abstract # 1169).

177OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Cook JC, Klinefelter GR, HardistyJF, Sharpe RM, and Foster PM (1999). Rodent Leydig cell tumorigenesis: A review of physiology, pathology, mechanisms and relevance to humans. Crit Rev Toxicol 29 , 169-261. Cope CS (1981). Measurement of Residual C-8 Fl uorosurfactant Levels on Various Products. U.S. Environmental Protection AgencyAdministrative Record 226-2533. Corsolini S and Kannan K (2004). Perfluorooctanesu lfonate and related fluorochemicals in several organisms including humans from Italy. Organohalogen Compounds 66 , 4079-4085. Cotran RS, Kumar V, and Robbins SL (1989). The pancreas. In: Pathologic Basis of Disease, pp. 981-1010. W.B. Saunders, Philadelphia. Crozier PW, Furdui V, Lucaciu CM, Stock NL, Mabur y SA, and Reiner EJ (2005) Determination of perfluorinated alkyl substances (PFAS) in sewage treatment plant effluents and biosolids by liquid chromatography-tandem mass spectrometry. Organohalogen Compounds 67 , 232-239 Cutts JH and Noble RL (1964). Estrone-induced mamm ary tumors in the rat. I. Induction and behavior of tumors. Cancer Res 24 , 1116-1123. D’Eon JC, Hurley MD, Wallington TJ, Mabury SA (2006). Atmospheric Chem istry of N- methyl Perfluorobutane Sulfonamidoethanol, C4F9SO2N(C H3)CH2CH2OH: Kinetics and Mechanism of Reaction with OH. Environ. Sci. Technol 40 (6) 1862-8 . Darby RT, Kaplan AM (1968). Fungal suscepti bility of polyurethanes. Appl. Microbiol 16 , 900- 905. Dai J, Li M, Jin Y, Saito N, Xu M and Wei F (2006). Perfluoroctanesulfonate and Perfluorooctanoate in Red Panda and Giant Panda from China. Environ Sci Technol (in press). Daikin (2000). Bioaccumulation Test of Perfluoroalkyl- carboxylic Acid (C= 7-13) in Carp. Test No. 51519, p. 26. Kurume Laboratory, Chemicals Eval uation and Research Institute, Japan (December 18). Daikin (2001). Chemicals Sales Co., LTD online Catalog. Fluoro aliptiatic compounds. http://daikin-dcs.co.jp/catalog/. De Silva AO and Mabury SA (2004). Isolating isom ers of perfluorocarboxylates in polar bears (Ursus maritimus) from two geographical locations. Environ Sci Technol 38 , 6538-6545. De Voogt P and van Roon A (2005). Perfluorinated ch emicals in sediments, particulate matter, soil, and water samples from The Neth erlands. Organohalogen Compounds 67 , 790-793 Dean WP and Jessup DC (1978). Acute oral toxicity(LD50) study in rats. International Research and Development Corporation, Study No. 137-091, May 5, 1978. U.S. Environmental Protection Agency Administrative Record 226-0419. Dinglasan MAJ, Ye Y, Edwards EA, Mabury SA (2004). Fluorotelomer Alcohol Biodegradation Yields Poly- and Perfluorinated Ac ids. Environ. Sci. Technol. 38, 2857-2864. Dinglasan MJ and Mabury S (2005). Evidence of 8:2 FTOH production from the biodegradation of 8:2 telomer methacrylate under aerobic conditions. Po ster presented at Fluoros: International Symposium on Fluorinated Alkyl Organic in the Environment, August 18-20, 2005. Toronto, Canada. Dinglasan-Panlilio MJ and Mabury SA (2006). Significant Residual Fl uorinated Alcohols Present in Various Fluorinated Materi als. Environ. Sci Technol 40 , 1447-1453.

178OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Dookhith, M. TY., 2001. Novel High Performance Water-Soluble Antifoam Based on Perfluoroalkylphosphinic and Perfluoroalkylphosphonic Acids in Crop Protection, Pesticide Formulations and Application Systems: 20th Vol. ASTM STP 1400. A.K. Viets, R.S. Tann and J. C. Mueninghoff Eds. ASTM, West Conshohcolen, PA. DuPont (1998). Meeting Notes and Charts for the January 6 and 7, 1998 PACE Team Meetings. U.S. Environmental Protection Agency Administrative Record 226-2084.
DuPont, 1999. RCRA Facility Investigation Report, DuPont Washington Works, West Virginia USEPA Permit Number WVD04-587-5291. Docket # OPPT-2003-0012-0184 . Dupont (Haskell Laboratory) (2001a). Letter to Ch arles Auer from Gerald Kennedy, dated January 25, 2001. FYI-00-001378. U.S. Environmental Prot ection Agency Administrative Record 226- 1000. Dupont (Haskell Laboratory) (2001b). Letter from Gerald Kennedy re: FYI-010-01378 and FYI0301-01378, dated May 29, 2001. U.S. Environmental Pr otection Agency Administrative Record 226-1023. DuPont (2002). Two-mile Radius Survey and C- 8 Sampling Report. U.S. Environmental Protection Agency Administrative Record 226-1235. December. Dupont (Haskell Laboratory) (200 3). Epidemiology surveillance report: Cancer incidence for Washington works site 1959-2001. U. S. Environmental Protection Agency Administrative Record 226-1307-6. DuPont (2003a). February 2003 Surface Water Monitoring Report for Washington Works Facility and Local, Letart and Dry Run Landfills; Washingt on, WV. U.S. Environmental Protection Agency Administrative Record 226-1508. April.DuPont (2003b). DuPont Telomer Manufacturing Sites: Environmental Assessment of PFOA Levels in Air and Water. U.S. E nvironmental Protection Agency Administrative Record 226-1534. September. DuPont (2003c). Sampling Investigation Results Little Hocking Water Association Well Field Washington County, Ohio. U.S. Environmental Pr otection Agency Administrative Record 226- 1509. April Dupont (2004). Letter from A. Michael Kaplan to 8(e) Coordinator re: 8EHQ-0381-0394, dated November 23, 2004. U.S. Environmental Protecti on Agency Administrative Record 226-1884. DuPont (2005). DuPont Global PFOA Strategy. U.S. EPA Administrati ve Record AR226-1914. U.S. Environmental Protecti on Agency, Washington, D. C. DuPont personel communiction (2005) E. van Wely, November 3 rd

. *DuPont (2005 a ). Company information for OECD Hazard Assessment Report. *DuPont (2005 b ). Personal communication from Andrew S. Hartten, Project Di rector, DuPont Corporate Remediation Group. Du Pont (2006). Quarterly MOU Status Report #2 Phase II Monitoring/Sampling Work Plan DuPont Washington Works (OPPT-2004-0113-0032 PFOA Site-Related Environmental Assessment Program). Dyneon (2004). Nutzen der Fluorkunststoffe für eine Nachhaltige Entwicklung. Presentation to the Germany Ministry of the Environment, Umwelt Bundes Amt (UBA), 30 November.

179OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Dyneon (2005). Personal communication M. Stubner Dyneon. December 14th. Eagon PK, Chander N, Epley MJ, Elm MS , Brady EP, and Rao KN (1994). Di(2- ethylhexyl)phthalate-induced chan ges in liver estyrogen metabolis m and hyperplasia. Int J Cancer 58 , 736-743. Edwards PJB, Jolley KW, Smith MH, Thomsen SJ, and Boden N (1997). Solven t isotope effect on the self-assembly liquid crystalline phase behavi or in aqueous solutions of ammonium pentadecafluorooctanoate. Langmuir 13 , 2665-2669. EG & G, Bionomics (1978). Report No. BW-78-6- 175, U.S. EPA Administra tive Record 226-0502, The Effects of Continuous Aqueous Exposure on Hatc hability of Eggs and Growth and Survival of Fry of Fathead Minnow, Pimephales promelas (June). Unpublished Data. Ellis DA, Mabury SA, Martin JW, and Muir DCG (2001). Thermolysis of fluoropolymers as a potential source of halogenated organi c acids in the environment. Nature 412 , 321-324.Ellis DA, Martin JW, De Silva AO, Mabury SA, Hurley MD, Sulbaek Andersen MP, and Wallington TJ (2004). Degradation of Fl uorotelomer Alcohols: A likely atmospheric sour ce of perfluorinated carboxylic acids. Environ Sci Technol 38 , 3316-3321. Emmett EA, Shofer FS, Zhang H, Freeman D, De sai C and Shaw LM (2006). Community Exposure to Perfluorooctanoate: Relationships between Seru m Concentrations and Exposure Scenarios. J Occup Med 48, 759-770. EPA (2004b). DuPont Telomer Manufacturing Sites: Environmental Assessment of PFOA Levels in Air and Water (September, 2003). USEP A Administrative Record AR226-2620 Eurofins (2005). Confidential study prot ocol 514021B/390-514025B/390 May 20 2005. Eurofins Analytik GmbH, Hamburg. Exfluor Research Company (1998). Material Safety Data Sheet for Pentadecafluorooctanoic acid, 96%, Acros Organics N.V. Online: http://216.122.170.58/msds .asp?ExfID=C8AC

. Fasano WJ, Kennedy GL, Szostek B, Farrar DG, Ward RJ, Haroun L, and Hinderliter PM (2005). Penetration of ammonium perfl uorooctanoate through rat and human skin in vitro. Drug Chem Toxicol 28, 79-90. Fasano WJ, Carpenter SC, Gannon SA, Snow TA , Stadler JC, Kennedy GL, Buck RC, Korzeniowski SH, Hinderliter PM, and Kemper RA ( 2006). Absorption, Distribution, Metabolism, and Elimination of 8:2 Fluorotelomer Alcohol inthe Rat. Toxicologi cal Sciences 91, 341-355. Fisher Scientific (2003). Mate rial Safety Data Sheet for C8 AC, Perfluorooctanoi c acid, 98%. Online: https://fscimage.fishersci.com/msds/97331.htm.FMG (2002). Fluoropolymer Manufacturing Group, “Fluoropolymer Manufacturers Group Presentation Slides.” U.S. EPA Administrative R ecord AR226-1094. U.S. Environmental Protection Agency, Washington, D. C. Frame SR and McConnell EE (2003). Review of prolif erative lesions of the exocrine pancreas in two chronic studies in rats with ammonium perfluorooctanoate. DuPont-13788. Oct. 16, 2003. Franklin J (2002). Screening Assessment for the Potential for Long-Range Atmospheric Transport of Perfluorooctanoic Acid. O PPT-2003-0012-0322. March 24, 2002. Furdui V, Crozier P, Reiner, E, and Mabury SA (2005). Direct measurement of perfluoralkylated surfactants in the Great Lakes wa ter samples. Organohalogen compounds 67 , 211-214

180OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Gabriel K (1976a). Primary eye irritation study in rabbits. Biosearch, Inc., March 4, 1976. U.S. Environmental Protection Agency Administrative Record 226-0422. Gabriel K (1976b). Primary skin irritation study in rabbits. Bi osearch, Inc., March 4, 1976. U.S. Environmental Protection AgencyAdministrative Record 226-0423. Gabriel K (1976c). Acute oral toxicity - rats . Biosearch, Inc., September 16, 1976. U.S. Environmental Protection Agency Administrative Record 226-0425. Gabriel K (1976d). Primary eye i rritation study in rabbits. Bios earch, Inc., September 16, 1976. U.S. Environmental Protection Agency Administrative Record 226-0426. Garry VF and Nelson RL (1981). An assay of ce ll transformation and cytotoxicity in C3H10T1⁄2 clonal cell line for the test ch emical T-2942 CoC. Stone ResearchLaboratories, Minneapolis, MN, March 4, 1981. U.S. Environmental Protecti on Agency Administrative Record 226-0428. Gauthier A and Mabury SA (2005) . Aqueous Photolysis of 8:2 Fluorotelomer Alcohol. Environ Toxicol Chem 24 (8), 1837-1846. Gazouli M, Yao ZX, Boujard N, Corton JC, Culty M, and Papadopoulos V (2002). Effect of peroxisome proliferators on Leydig cell periferal-type benzodiazepine receptor gene expression, hormone-stimulated chloresterol transport, and steroidogenesis: Role of the proxisome proliferatoractivated receptor α . Endocrinology 143 , 2571-2583. Gibson SJ and Johnson JD (1979). Ab sorption of FC-143-14C in rats after a single oral dose. Riker Laboratories, Inc., Subsidiaryof 3M, St. Paul, Minnesota. Gibson SJ and Johnson JD (1980). Extent and route ofexcretion and tissue distribution of total carbon-14 in male and female rats after a single I. V. dose of FC-143-14C. Riker Laboratories, Inc., Subsidiary of 3M, St. Paul, MN. Gibson SJ and Johnson JD (1983). Extent and route of excretion of to tal carbon-14 in pregnant rats after a single oral dose of amm onium 14 C-perfluorooctanoate. Riker Laboratories, Inc., Subsidiary of 3M, St. Paul, MN. Giesy JP and Kannan K (2001). Global distribution of perfluorooctane sulfona te in wildlife. Environ Sci Technol 35 , 1339-1342. Gilliland F (1992). Fluorochemicals and human health: Studies in an occupational cohort. Doctoral thesis, Division of Environmental and Occupationa l Health, University of Minnesota. U.S. Environmental Protection Agency Ad ministrative Record 226-0473. Gilliland FD, Mandel JS (1993). Mortality among empl oyees of a perfluorooctanoic acid production plant. J Occup Med 35 , 950-954. Gilliland FD, Mandel JS (1996). Serum perfluorooctanoic acid and hepatic enzymes, lipoproteins, and cholesterol: A study of occupatio nally exposed men. Am J Ind Med 29 , 560-568. Glaza S (1995). Acute dermal toxici ty study of T-6342 in rabbits. Co rning Hazelton, Inc. Madison, WI. Project ID: HWI 50800374. 3M Company. St. Pa ul, MN. U.S. Environmental Protection Agency Administrative Record 226-0427. Glaza SM (1997). Acute oral toxicity study of T- 6669 in rats. Corning Hazleton Inc. Study No. CHW 61001760, January 10, 1997. U.S. Environmental Protection Agency Administrative Record 226-0420.

181OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Gobas, F.A.C.P. and Morrison, H.A. (2000). Bioc oncentration and biomagni fication in the aquatic environment. Chapter 9 In: Boethling, R.S. and Mackay, D. (e ds.). Handbook of property estimation methods for chemicals, environmental and hea lth sciences. CRC Press, Boca Raton, USA. pp. 189-231 Goldenthal E.I. (1978a). Final Report, NinetyDay Subacute Rat T oxicity Study on Fluorad® Fluorochemical FC-143, Internati onal Research and Development Corporation, Study No. 137-089, 3M Reference No. T-3141, Novemb er 6, 1978. U.S. Environmental Protection Agency Administrative Record 226-0441. Goldenthal E.I. (1978b). Final Report, Ninety Day Subacute Rhesus Monkey Toxicity Study, International Research and Development Corpor ation, Study No. 137-090, November 10, 1978. U.S. Environmental Protection AgencyAdministrative Record 226-0447.
Goss K-U, Bronner G, Harner T, Hertel M, and Schmidt TC ( 2006). The Partition Behavior of Fluorotelomer Alcohols and Olefins. Environ. Sci. Technol. 40, 3572-3577. Gortner EG (1981). Oral teratol ogy study of T-2998CoC in rats. Safety Evaluation Laboratory and Riker Laboratories, Inc. Experi ment Number: 0681TR0110, December 1981. Gortner EG (1982). Oral terato logy study of T-3141CoC in rabbits . Safety Evaluation Laboratory and Riker Laboratories, Inc. Expe riment Number: 0681TB0398, February 1982. Gray, JS. 2002. Biomagnification in marine systems: the perspective of an ecologist. Mar. Poll. Bull. 45: 46-52. Griffith FD and Long JE (1980). Animal toxicitystudies with ammonium perfluorooctanoate. Am Ind Hyg Assoc J 41 , 576-583. Guruge KS, Taniyasu S, Miyazaki S, Yamanaka N, and Yamashita N (2004). Age dependent accumulation of perfluorinated chemicals in beef cattles. Organohalogen Compounds 66 , 4029-4034. Guruge KS, Taniyasu S, Yamashita N, Wijeratn a S, Mohotti KM, Seneviratne HR, Kannan K, Yamanaka N, and Miyazaki S (2005 ). Perfluorinated organic co mpounds in human blood serum and seminal plasma: a study of urban and rural tea wo rker populations in Sri Lanka. J Environ Monit 7 , 371-377. Guruge K, Manage PM, Miyazaki S, Yamanaka N, Taniyasu S, Hanari N, and Yamashita N (2005a). Species-specific accumulation of perfluorin ated compounds in farm animals in Japan. Organohalogen Compounds 67 , 823-826 Han X (2003) Ammonium perfluorooctanoate: Age effect on the plasma concentration in postwealning rats following oral gava ge. Haskell Laboratory for Healthand Environmental Sciences. Study No. Dupont-13267, December 15, 2003 U.S. Enviro nmental Protection Agency Administrative Record 226-1553. Han H, Snow TA, Kemper RA, and Jepson GW ( 2003). Binding of perfluor ooctanoic acid to rat and human plasma proteins. Chem Res Toxicol 16 , 775-781. Hanhijarvi H, Ophaug RH, and Singer L (1982). Th e sex-related difference in perfluorooctanoate excretion in the rat. Proc Soc Exp Biol Med 171 , 50-55. Hanhijarvi H, Ylinen M, Kojo A, and Kosma VM (1987). Elimination and toxicity of perfluorooctanoic acid during subchronic administration in the Wistar rat. Pharmacol Toxicol 61 , 66-68.

182OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Hanhijarvi H, Ylinen M, Haaran en T, and Nevalainen T (1988). A proposed species difference in the renal excretion of perfluorooc tanoic acid in the beagle dog and rat. In: New Developments in Biosciences: Their Implications for Laboratory Animal Science. Beynen AC, Solleveld HA, Eds. Martinus Nijhoff Publishers. Dordrecht, The Netherlands. Hansch C and Leo A (1979). The fragment method of calculated partition coefficients. In: Substituent Constants for Correlation Analysis and Chem istry and Biology, Chapter I.V. John Wiley and Sons, Inc., New York. Hansen KJ, Clemen LA, Ellefson ME, and J ohnson HO (2001). Compound-specific, qualitative characterization of organic fluorochemicals in biological matrices. Environ Sci Technol 35 , 766- 770. Hansen KJ, Johnson HO, Eldridge JS, Butenhoff JL , and Dick LA (2002). Quantitative characterization of trace levels of PFOS and PFOA inthe Tennessee River. Environ Sci Technol 36 , 1681- 1685. Hanson ML, Small J, Sibley PK, Boudreau TM, Brain RA, Mabury SA, a nd Solomon KR (2005). Microcosm evaluation of the fate , toxicity, and risk to aquatic macrophytes from perfluorooctanoic acid (PFOA). Arch Environ Contam Toxicol 49 , 307-316. Harada K, Saito N, Inoue K, Yoshinaga T, Watanabe T, Sasaki S, Kamiyama S, and Koizumi A (2004). The influence of time, sex, and geographic f actors on levels of perfl uoroctane sulfonate and perfluorooctanoate in human serum over the last 25 years. J Occup Health 46 , 141-147. Harada K, Inoue K, Morikawa A, Yoshinaga T, Saito N, and Koizumi A ( 2005). Renal clearance of perfluorooctane sulfonate and perf luorooctanoate in humans and th eir species-specific excretion. Environ Res 99 , 253-261. Harada K, Nakanishi S, Saito N, Tsutsui T, and Koizumi A (2005). Air borne perfluorooctanoate may be a substantial source contamination in Kyoto area, Japan. Bull Environ Contam Toxicol 74 , 64-69. Hardisty JF (2005). PWG report: Pathology peer re view and pathology wo rking group review of mammary glands from a chronic feeding study in rats with PFOA. Experimental Pathology Laboratories, Inc. Research Triangle Park, NC. June 17, 2005. Higgins CP, Field JA, Criddle CS, and Luthy RG (2005). Quantitative determination of perfluorochemicals in sediments and dome stic sludge. Environ Sci Technol 39 , 3946-3956. Hinderliter PM (2003). Perfluoroocta noic acid: relationship between repeated inhalation exposures and plasma PFOA concentration in the rat. Hask ell Laboratory for Health and Environmental Sciences. Study No. DuPont-12944, November 5, 2003. Hinderliter PM (2004). Ammonium perfluorooctanoate: Age effect on the PFOA plasma concentration in post-weaning rats following oral gavage. E.I. du Pont de Nemours and Company. Laboratory Project ID: Dupont-15302. December 2, 2004. Hinderliter PM, Mylchreest E, Gannon SA, Butenhoff JL, and Kennedy GL Jr (2005) Perfluorooctanoate: Placental and lact ational transport pharmacoki netics in rats. Toxicology 211, 139-48. Hoff PT, Van de Vijver K, Van Dongen W Esma ns EL, Blust R, and De Coen WM (2003). Perfluorooctane sulfonic acid in bib (Trisopterus luscus) and plaice (P leuronectes platessa) from the Western Scheldt and the Belgian No rth Sea: distribution and bioche mical effects. Environ Toxicol Chem 22 , 608-614

183OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Hoff PT, Scheirs J, Van de Vijver K, Van Dongen W, Esmans EL, Blust R, and De Coen W (2004). Biochemical effect evaluation of perfluorooctane sulfonic acid–contaminated wood mice (Apodemus sylvaticus). Environ Health Perspect 112 , 681–686. Hohenblum P, Scharf S, and Sitka A (2003). Perfl uorierte anionische Tenside in österreichischen Industrieabwässern. Vom Wasser 101 , 155-164. Holmström K, Järnberg U, and Bi gnert A (2005). Temporal trends of PFOS and PFOA in guillemot eggs from the Baltic Sea, 1968-2003. Environ Sci Technol 39 , 80-84. Hori H, Hayakawa E, Einaga H, Kutsuna S, Koike K, Ibusuki T, Kiatagawa H, and Arakawa R(2004a). Decomposition of environmentally persistent perfluorooctanoic acid in water by photochemical approaches. Environ Sci Technol 38 , 6118-6124. Hori H, Hayakawa E, Yamashita N, TaniyasuS, Nakata F, and Kobayashi Y (2004b). Highperformance liquid chromatography with conductimetric detection of perfluorocarboxy lic acids and perfluorosulfonates. Chemosphere 57 , 273-282. Hori H, Yamamoto A, Hayakawa E, Taniyasu S, Yamashita N, and Kutsuna S (2005). Efficient decomposition of environmentally persistent perfluor ocarboxylic acids by use of persulfate as a photochemical oxidant. Environ. Sci. Technol. 39 , 2383-2388. Houde M, Wells RS, Fair PA, Bossart GD, HohnAA, Rowles TK, Sweeney JC, Solomon KR, and Muir DCG (2005) Perfluoroalkyl compounds in fr ee-ranging bottlenose dol phins (Tursiops truncates) from the Gulf of Mexico and th e Atlantic Ocean. E nviron Sci Technol 39 , 6591-6598. Houde M, Bujas TAD, Small J, Wells RS, Fair PA, Bossart GD, Solomon KR and Muir DCG (2006). Biomagnification of Perfl uoroalkyl Compounds in the Bottle nose Dolphin (Tursiops truncatus) Food Web. Environ Sci Technol 40 , 4138-4144. Howard GT (2002). Biodegradation of polyurethane : a review. Internati onal Biodeterioration & Biodegradation, 49 , 245-252. Hurley MD, Sulbaek Andersen MP, Wallington TJ , Ellis DA, Martin JW, and Mabury SA (2004a) Atmospheric chemistry of perfluorinated carboxy lic acids: reaction with OH radicals and atmospheric lifetimes. J. Phys. Chem. A 2004, 108, 615-620. Hurley MD, Wallington TJ, Sulbaek Anderson MP, Ellis DA, Martin JW, and Mabury SA (2004b). Atmospheric chemistry of fluorinated alcohols: r eaction with Cl atoms and OH radicals and atmospheric lifetimes. J. Phys. Chem. A.. Vol 108 : 1973 – 1979 IARC (International Agency for Research on Can cer) (1995). WHO Interna tional Agency for Research on Cancer. Peroxisome proliferation and its role in carcinogenesis. IARC technical report No. 24. IARC Press, Lyon, France. Ikeda T, Aiba K, Fukuda K, and Tanaka M (1985). The induction of peroxisome proliferation in rat liver by perfluorinated fatty acids, metabolically inert derivatives of fatty acids. J Biochem 98 , 475- 482. Inoue K, Okada F, Ito R, Kato S, Sasaki S, Na kajima S, Uno A, Saijo Y, and Sata F (2004). Perfluorooctane sulfonate (PFOS) a nd related perfluorinated compounds in human maternal and cord blood samples: assessment of PFOS exposure in a susceptible population during pregnancy. Environ Hlth Persp 112 , 1204-1207.

184OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Johnson JD, Gibson SJ, and Ober RE (1984). Cholestyramine-enhan ced fecal elimination of carbon-14 in rats after admini stration of ammonium [ 14 C]perfluorooctanoate or potassium [ 14 C]perfluorooctanesulfonate . Fundam Appl Toxicol 4 , 972-976. Johnson JD (1995a). Final report, analytical study, single-dose in travenous pharmacokinetic study of T-6067 in rabbits. Study Nu mber: AMDT-120694.1. 3M Environmental Technology & Services, St. Paul, MN. U.S. Environmental Protecti on Agency Administrative Record 226-0453. Johnson JD (1995b). Final report, an alytical study, single-dose abso rption/toxicity study of T-6067, T-6068, and T-6069 in rabbits. St udy Number: AMDT-011095.1. 3M Environmental Technology & Services, St. Paul, MN. U.S. Environmental Pr otection Agency Administrative Record 226-0454. Johnson B, Olivero J, and Kannan K (2004). Perfl uorinated compounds in fish, birds and humans from Colombia. Society of Environmental Toxi cology and Chemistry North America Annual Meeting. Abstract PW177. Kachanova ZP and Koslov JN (1973). Zh. Fiz. Khim. 47 , 2107. Kaiser MA, Cobranchi DP, Kao C-PC, Krusic PJ , Marchione AA., and Buck RC. (2004). Physicochemical Properties of 8:2 Fluorinated Telomer B Alcohol. Journal of Ch emical and Engineering Data 49, 912-916. Kaiser MA, Larsen BS, Kao CPC, and Buck RC (2005). Vapor Pres sures of Perfluorooctanoic, - nonanoic, -decanoic, -undecanoic, and -dodecanoic Ac ids. Journal of Chemical and Engineering Data 50 , 1841-1843. Kaiser MA, Larsen BS, Dawson BJ, Kurtz K, Lieckfi eld R jr, Miller JR, a nd Flaherty J (2005a). Methof for the determination of perfluorooctanoic acid in air samples using liquid chromatography with mass spectrometry. J Occup Environ Hyg 2 , 307-313 Kallenborn R, Berger U, and Järnbe rg U (2004). Perfluorinated alky lated substances (PFAS) in the Nordic Environment. Nordic Council of Ministers, Tromso, Norway. Kannan K, Choi JW, Iseki N, Senthilkumar K, Kim DH, Masunaga S, and Giesy JP (2002a). Concentrations of perfluorinated acids in livers of birds from Japan and Korea. Chemosphere 49 , 225- 231. Kannan K, Corsolini S, Falandysz J, Oehme G, Fo cardi S, and Giesy JP (2002b) Perfluorooctanesulfonate and related fluorinated hydrocarbons in marine mammals, fishes, and birds from coasts of the Baltic and the Mediterrane an Seas. Environ Sci Technol 36 , 3210-3216. Kannan K, Newsted J, Halbrook, RS, and Giesy JP (2002c). Perfluorooctanes ulfonate and related fluorinated hydrocarbons in mink and river otters from the United States. Environ Sci Technol 36 , 2566-2571. Kannan K, Corsolini S, Falandysz, J, Fillmann G, Kumar KS, Loganathan BF, Mohd M, Olivero J, Van Wouwe N, Yang JH, and Aldous KM (2004). Perfluorooctanesulfonate and related fluorochemicals in human blood from seve ral countries. Environ Sci Technol 38 , 4489-4495. Kannan K, Tao L, Sinclair E, Pa stva SD, Jude DJ, and Giesy JP (2005). Perfluorinated compounds in aquatic organisms at various trophic levels in a Great Lakes food chain. Arch Environ Contam Toxicol 48 , 559-566. Kannan, K., Perotta, E., and Thomas, N.J. (2006) Association between Perfluorinated Compounds and Pathological Conditions in Southern S ea Otters. Environmental Science and Technology 40: 4943-4948

185OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Karns ME and Fayerweather WE (1991). A case-co ntrol study of leukemia at the Washington Works site. Final Report. Dupont Company. De cember 31, 1991. U.S. Envi ronmental Protection Agency Administrative Record 226-1308-2. Kawashima Y, Uy-Yu N, and Kozuka H (1989). Se x-related difference in the inductions by perfluorooctanoic acid of peroxisome α -oxidation, microsomal 1-acy lglycerolphosphocholine acyltransferase and cytosolic long-chain acyl -CoA hydrolase in ra t liver. Biochem J 261 , 595-600. Kauck EA and Diesslin AR (1951) . Some properties of perfluoroc arboxylic acids. Ind Eng Chem 43 , 2332-2334. Keller BJ, Marsman DS, Popp JA, and Thurman RG (1992). Several nongenotoxic carcinogens uncouple mitrochondrial oxidative phos phorylation. Biochem Biophys Acta 1102 , 237-244. Keller J, Kannan K, Taniyasu S, Yamashita N, Day R, Arendt M, Maier P, Segars A, Whitacker J, and Kucklick J (2004). Perfluorinat ed organic compounds in the plas ma of juvenile loggerhead and Kemp’s ridley sea turtles from the southeastern coast of the U.S. Society of Environmental Toxicology and Chemistry North America Annua l Meeting, Portland, Oregon. Abstract 810. Keller JM, Kannan K, Taniyasu S, Yamashita N, Day RD, Arendt MD, Segars AL, and Kucklick JR (2005). Perfluorinated compounds in the plasma of loggerhead and Kemp’s ridley sea turtles from the southeastern coast of the United States. Environ Sci Technol 39 , 9101-9108. Kelly, B.C., F.A.P.C. Gobas and M.S. McLachlan. (2004). Intestinal absorp tion and biomagnification of organic contaminants in fish, w ildlife and humans. Environ. Toxicol. Chem. 23 (10): 2324- 2336. Kelly JW, Purkey HE, Palaninathan SK, Kent KC , Smith C, Safe SH, and Sacchettini JC (2004). Hydroxylated polychlorinated biphe nyls selectively bind transthyretinin blood and inhibit amyloidogenesis: rationalizing r odent PCB toxicity. Chem istry and Biology 11: 1719-1728. Kemper RA (2003). Perfluorooctanoic acid: Toxicokinetics in the ra t. Laboratory Project ID: Dupont-7473. Haskell Laboratory for Health and Envi ronmental Sciences, E.I. du Pont de Nemours and Company. April 2, 2003. U.S. Environmental Pr otection Agency Admini strative Record 226- 1499. Kennedy GL (1985). Dermal toxicity of ammonium perfluorooctanoate. Toxicol Appl Pharmacol 81 , 348-355. Kennedy GL, Hall GT, Brittelli MR, Barnes JR, and Chen HC (1986). Inhalation toxicity of ammonium perfluorooctanoate. Food Chem Toxicol 24 , 1325-1329. Kerstner-Wood C, Coward L, and Gorman G (2003). Protein binding of perfluorohexane sulfonate, perfluorooctane sulfonate and perf luorooctanoate to plasma (human, rat, and monkey), and various human-derived plasma protein fr actions. Southern Research Inst itute. Study ID 9921.7. U.S. Environmental Protection Agency Ad ministrative Record 226-1354. Kissa E (2001). Fluorinated Surfactants and Re pellents, Marcel Dekker, New York, Vol. 97. Klasmeier J, Wegmann F and Matthies M (2004) Ev aluation of Environmental Exposure and Longrange Transport Potential for Perfluoronated 8:2 Telomer B Alcohol. USEPA Administrative Record AR226-3359. Klaunig JE, Babich MA, Baetcke KP, Cook JC , Corton JC, David RM, DeLuca JG, Lai DY, McKee RH, Peters JM, Roberts RA , and Fenner-Crisp PA (2003). PPAR α agonist-induced rodent tumors: Mode of action and human relevance. Crit Rev Toxicol 33 , 655-780.

186OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Krusic PJ and Roe DC (2004). Gas-phase NMR technique for studying the ther molysis of materials: thermal decomposition of ammonium perf luorooctanoate. Analytical Chemistry 76 (13), 3800- 3803.Krusic PJ, Marchione AA, Davidson F, Kaiser MA, Kao C-PC, Richardson RE, Botelho M, Waterland RL, and Buck, RC (2005). Vapor Pressure and Intramolecular Hydrogen Bonding in Fluorotelomer Alcohols. Journal of Physical Chemistry A 109, 6232-6241. Kubwabo C, Vais N, and Benoit FM (2004). A pilo t study on the determination of perfluorooctanesulfonate and other perfluor inated compounds in blood of Canadians. J Environ Monit 6 , 540-545. Kubwabo C, Stewart B, Zhu J, and Marro L (2 005). Occurrence of perfluor osulfonates and other perfluorochemicals in dust from selected homes in the city of Ottawa, Canada. J Environ Monit 7 ,1074-1078. Kudo N, Katakura M, Sato Y, and Kawashima Y (2002). Sex hormone-regulated renal transport of perfluorooctanoic acid. Chem Biol Interact 139 , 301-316. Kuklenyik Z, Reich JA, Tully JS, Needham LL, and Calafat AM (2004). Automated solid-phase extraction and measurement of perfluorinated organic acids and amides in human serum and milk. Environ Sci Technol 38 , 3698-3704. Kurunthachalam S, Kurunthachalam K, Ohi E, Koizumi A, and Takasuga T (2005) Occurrence ofperfluorinated contaminants in water, sediment and fish from Kyoto area, Japan. Organohalogen 67 , 229-231. Lange, CC (2000) The aerobic biodegradation of N- EtFOSE alcohol by the microbial activity present in municipal wastewater treatment sludge. P ace Analytical Services, Science Solutions Division, Minneapolis, MN project CA058 for 3M Environmental Laboratory. November 2, 2000. Lange CC (2002). Biodegradation Sc reen Study for Telomer-Type Al cohols. 3M Project ID No. E01-0684. US EPA Public Docket ARR-226-0555. Lange FT, Schmidt CK, Metzinger M, Wenz M, and Brauch HJ (2004). Determination of perfluorinated carboxylates and sulfona tes from aqueous samples by HPLC-ESI-MS-MS and their occurrence in surface waters in Germany. Society of Environmental Toxicology and Chemistry Europe Annual Meeting, Prague. Poster MOPO15/006. 0Larsen BS, Kaiser MA, Botelho M, Wooler GR , and Buxton LW (2005). Comparison of pressurized solvent and reflux extraction methods for the determination of pe rfluorooctanoic acid in polytetrafluoroethylene polym ers using LC-MS-MS. Analyst 130 , 59-62. Larsen, B. S., Stchur, P., Szoste k, B., Bachmura, S. F., Rowand, R. C., Prickett, K. B., Korzeniowski, S. H., and Buck, R. C. (2006). Method development for the determination of residual fluorotelomer raw materials and perflurooctanoate in fluorotelomer-based products by gas chromatography and liquid chromatograp hy mass spectrometry. Journal of Chromatography, A 1110, 117- 124. Lau C, Thibodeaux JR, Hanson RG, Narotsky MG, Rogers JM, Lindstrom AB, and Strynar MJ (2005). Effects of Perfluorooctoanoic Acid Exposure During Pregnancy in the Mouse. Tox Sci (in press). Lawlor TE (1995). Mutagenicity test with T-6342 in the Salmonella-Escherichia coli /mammalianmicrosome reverse mutation assay. Laboratory Number: 17073-0-409. Corni ng Hazleton Inc., Vienna, VA. 3M Company. St. Paul, MN. U.S. Envi ronmental Protection Agency Administrative Record 226-0436.

187OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Lawlor TE (1996). Mutagenicitytest with T-6564 in the Salmonella–Escherichia coli/mammalianmicrosome reverse mutation assay with a confirmato ry assay. Corning Hazlet on Inc. Final Report. CHV Study No: 17750-0-409R. September 13, 1996. U.S. Environmental Protection Agency Administrative Record 226-0432. Leung L, Kasprzak DJ and English AD (2002). DuPont Corporate Center for Analytical Sciences: PTFE and FEP Thermal Stability. U.S. Environm ental Protection Agency Administrative Record 226-2685. Li Y, Demerjian KL, Williams LR, Worsnop DR, Kolb CE, and Davidovits P (2006). Heterogeneous Uptake of 8:2 Fluorotelom er Alcohol on Liquid Water and 1- Octanol Droplets. Journal of Physical Chemistry A 110, 6814-6820. Lide DR, editor (2003). CRC Handbook of Chemistr y and Physics. 84th edition. CRC Press. Lines D and Sutcliffe H (1984). Prep aration and properties of some salts of perfluorooctanoic acid. J Fluorine Chem 25 , 505-512. Liu RCM, Hurtt ME, Cook JC, and Biegel LB (1996). Effect of the peroxisome proliferator, ammonium perfluorooctanate (C8), on hepatic aromatas e activity in adult male Crl:CD BR (CD) rats. Fund Appl Toxicol 30 , 220-228. Liu SC, Sanfilippo B, Perroteau I, Derynck R, Salomon, DS, and Kidwell WR (1987). Expression of transforming growth factor alpha (TGF alpha) in differentiated rat mammary tumors: estrogen induction of TGF alpha production. Mol Endocrinol 1 , 683-692. Liu J, Nies, Nakatsu L, Lee CL (2 005). Sorption and biotransforma tion of 8:2 fluorotelomer alcohol by surface soils. Poster presented at SETAC Mee ting, Perfluorochemical Bi odegradation, Fate, and Transport. November 17, 2005. Baltimore, U.S.A. Loewen M, Halldorson T, Wang F, and Tomy G (2005) . Fluorotelomer carboxylic acids and PFOS in rainwater from an urban cente r in Canada. Environ Sci Technol 39 , 2944-2951. Longnecker DS (1987). Interface be tween adaptive and neoplastic growth in the pancreas. Gut 28 , 253-258. Longnecker DS and Sumi C (1990). Effects of sex steroid hormones on pancreatic cance r in the rat. Int J Pancreatol 7 , 159-165. Loveless SE, Finlay C, Everds NE, Frame SR, Gillies PJ, O’Connor JC, and Kennedy GL (2006). Comparative responses of rats andmice exposed to linear/branched, linear, or branched ammonium perfluorooctanoate (APFO). Toxicology 220 , 203-217. Lucaciu C, Furdui VI, Crozier PW, Reiner EJ, Marvin CH, Wania F, and Mabury SA (2005). Temporal study of perfluorinated al kyl surfactants in Niagara rive r sediments (1980-2002). Organohalogen Compounds 67 , 764-766 Luebker DJ, Hansen KJ, Bass NM, Butenhoff JL, and Seacat AM (2002). Interactions of fluorochemicals with rat liver fatty acid-binding proteins. Toxicology 176 , 175-185. Lunak S and Sediak P (1992). Photoinitiated reactio ns of hydrogen peroxide in the liquid phase. J. Photochem. Photobiol. A. Chem. 68 , 1-33.
Mackintosh, C, J Moldonado, JH Wu, N Hoover, A Chong, MG Ikonomou, and FAPC Gobas. (2004). Distribution of phthalate esters in a marine aquatic food web: comparison to polychlorinated biphenyls. Environ. Sci. Technol. 38:2011-2020.

188OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Mackay, D and A Fraser. 2000. Bioaccumulation of pe rsistent organic chemicals: mechanisms and models. Environmental Pollution 110: 375-391. Mackay D (2001). Multimedia environmental m odels: The Fugacity Approach. 2nd Edition. CRC Press, Boca Raton, Florida, USA. Maloney EK and Waxman DJ (1999). Transactivation of PPAR α and PPAR γby structurally diverse environmental chemicals. Toxicol Appl Pharmacol 161 , 209-218. Mann PC and Frame SR (2004). FC-143: Two year or al toxicity-oncogenicit y study in rats: Peer review of ovaries. DuPont Pr oject ID 15261, June 25, 2004. U.S. Environmental Protection Agency Administrative Record 226. Mann PC, Boorman GA, Lollini LO, McMartin DN, and Goodman DG (1996). Proliferative lesions of the mammary gland in rats. IS-2. In: Guides for Toxicologic Pathol ogy. STP/ARP/AFIP, Washington, D.C.. Markoe DM (1983). Primary skin irritation test with T-3371 in albino rabbits. Riker Laboratories, Study No.0883EB0079, July 13, 1983. U.S. Environmen tal Protection Agency Administrative Record 226-0424. Martin JW, Muir DCG, Moody CA, Ellis DA, Kwan WC, Solomon KR, Mabury SA (2002). Collection of Airborne Fluorinated Organics and An alysis of Gas Chromatography / Chemical Ionization Mass Spectrometry. Anal Chem 74 , 584-590. Martin JW, Mabury SA, Solomon KR, and Muir DC G (2003a). Dietary accu mulation of perfluorinated acids in juvenile rainbow trout ( Oncorhynchus mykiss ). Environ Toxicol Chem 22 , 189-195. Martin JW, Mabury SA, Solomon KR, and Muir DCG (2003b). Bioconcentration and tissue distribution of perfluorinated acids in rainbow tr out (Oncorhynchus mykiss). Environ Toxicol Chem 22 , 196-204. Martin JW, Whittle DM, Muir DC, and Mabury SA ( 2004a). Perfluoroalkyl contaminants in a food web from Lake Ontario. Environ Sci Technol 38 , 5379-5385. Martin JW, Smithwick MM, Braune BM, HoekstraPF, Muir DCG, and Mabury SA (2004b). Identification of long-chain perfluorinated acids in biota from the Cana dian Arctic. Environ Sci Technol 38 , 373-380. Martin J, Kannan K, Berger U, de Voogt P, Field J, Fr anklin J, Giesy JP, Ha rner T, Muir CGD, Scott B, Kaiser MA, Jaernberg U, Jones KC, Mabury SA, Schroeder H, Simcik M, Sottani C, Van Bavel B, Kaerrman A, Lindstroem G, van Leeuwe n S (2004c). Analytical challenges hamper perfluoroalkyl research. Envi ron Sci Technol 38, 248A-255A. Martin JW, Ellis DA, Mabury SA, Hurley MD and Wallington TJ (2006). Atmospheric Chemistry of Perfluoroalkanesulfonamides: Kinetic and Produc t Studies of the OH Radical and Cl Atom Initiated Oxidation of N-Ethyl Perfluorobutanes ulfonamide. Environ. Sci. Technol 40(3): 864-872. Mawn MP, McKay RG, Ryan TW, Szostek B, Powl ey CR, and Buck RC (2005). Determination of extractable perfluorooctanoic acid (PFOA) in water, sweat simulant , saliva simulant, and methanol from textile and carpet samples by LC/MS/MS. Analyst 130 , 670-678. McDonald M, Warne AL, Stock NL, Mabury SA, So lomon KR, and Sibley PK (2004). Toxicity of perfluorooctane sulfonic acid and perfluorooctanoic acid to Chironom us tentans. Environ Toxicol Chem 23 (9), 2116-2123.

189OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

MDH (Minnesota Department of Health) (2004). Ground-water sampling for PFOA near the Washington Co. Landfill. Available from www.health.state.mn.us, 3 p.Meesters RJW and Schroeder HF (2004). Perfluorooctane sulfonate – a quite mobile anionic anthropogenic surfactant, ubiquitously found in the environment. Water Sci Technol 50 , 235-242. Merck KGaA (2005) Material Safety Data Sheet January 20 Metrick M and Marias AJ (1977). 28-Day Oral Toxicity Study with FC-143 in Albino Rats, Final Report, Industrial Bio-Test Laboratories, Inc. Study No. 8532-10654, 3M Reference No. T1742CoC, Lot 269, September 29, 1977. Miteni (2005). Personal communication from F. Esposito, Miteni SpA. October 27 th

. Moody CA and Field JA (1999). Determination of perfluorocarboxylates ingroundwater impacted by fire-fighting activity. Environ Sci Technol 33 , 2800-2806. Moody CA, Kwan WC, Martin J, Muir DCG, and Mabury SA (2001). Determination of perfluorinated surfactants in surface water samples by two different analyti cal techniques: liquid chromatography/tandem mass spectrometry and 19F NMR. Anal Chem 73 , 2200-2206. Moody CA, Martin JW, Kwan WC, Muir DC, and Mabury SA (2002) . Monitoring perfluorinated surfactants in biota and surface water samples follo wing an accidental release of fire-fighting foam into Etobicoke Creek. Environ Sci Technol 36 , 545-551. Moody CA, Hebert GN, Strauss SH, and Field JA(2003). Occurrence an d persistence of perfluorooctanesulfonate and other perf luorinated surfactants in groundwat er at a fire-training area at Wurtsmith Air Force Base, Michigan, USA. J Environ Monit 5 , 341-345.Morikawa A, Kamei N, Harada K, Inoue K, Yoshinaga T, Saito N, and Koizumi A (2005) The bioconcentration factor of perfluorooc tane sulfonate is significantly larger than that of perfluorooctanoate in wild turtles (Trachemys scripta eleg ans and Chimneys reevesii) An Ai river ecological study. Ecotox Environ Safety in press. Moriwaki H, Takata Y, and Ar akawa R (2003). Concentrations of perfluorooctane sulfonate (PFOS) and perfluorooctanoic acid (PFOA) in vac uum cleaner dust collected in Japanese homes. J Environ Monit 5 , 753-757. Moroi Y, Yano H, Shibata O, and Yonemitsu T ( 2001). Determination of acidity constants of perfluoroalkanoic acids. Bull Chem Soc Japan, 74 (4), 667-672. Murli H (1995). Mutagenicity test on T-6342 in anin vivo mouse micronucleus assay. Corning Hazleton Inc., Vienna, VA. Study No. 17073-0-455, December 14, 1995. U.S. Environmental Protection Agency Administrative Record 226-0435. Murli H (1996a). Mutagenicity test on T-6564 in an in vivo m ouse micronucleus assay. Corning Hazleton Inc., Vienna, VA. Study number 17750-0-455, November 1, 1996. U.S. Environmental Protection Agency Admini strative Record 226-0430. Murli H (1996b). Mutagenicity test on T-6564 meas uring chromosomal aberrations in Chinese Hamster Ovary (CHO) cells with a confirmatory assay with multiple harvests. Corning Hazleton Inc., Vienna, VA. Study No. 17750-0-437CO, Sept ember 16, 1996. U.S. Environmental Protection Agency Administrative Record 226-0431. Murli H (1996c). Mutagenicity test on T-6342 meas uring chromosomal aberrations in human whole blood lymphocytes with a confirmatory assay withmultiple harvests. Corning-Hazelton, Inc., Vi190OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

enna, VA. Study No. 17073-0-449CO, November 1, 1996. U.S. Environmental Protection Agency Administrative Record 226-0433. Murli H (1996d). Mutagenicity test on T-6342 meas uring chromosomal aberrations in Chinese Hamster Ovary (CHO) cells with a confirmatory assay with multiple harvests. Corning-Hazelton, Inc. Vienna, VA. Study No. 17073-0-437CO, Sept ember 16, 1996. U.S. Environmental Protection Agency Administrative Record 226-0434. Mylchreest E (2003). PFOA: Lactational and pl acental transport pharmacokinetic study in rats. Haskell Laboratory for Health and Environmen tal Sciences, Newark DE, Study No. DuPont-13309, December 19, 2003. U.S. Environmental Protecti on Agency Administrative Record 226-1551. Nakajima-Kambe, T, Shigeno-Akut su Y, Nomura N, Onuma F, Na kahara, T (1999). Microbial degradation of polyurethane, polyester polyurethanes and polyether polyurethane s, Appl Microbiol and Biotech 51 , 134-140. Nakata H, Kannan K, Nasu T, Cho H-S, SinclairE, and Takemura A (2005) Perfluorinated contaminants in tidal flat and shallow water areas from the Ariake Sea, Japan. Organohalogen Compounds 67 , 753-755 Nilsson R, Beije B, Preat V, Erxon K, and Rame l C (1991). On the mechanism of the hepatocarcinogenicity of peroxisome prolif erators. Chem Biol Interact 78 :235-250. NIOSH (National Institute of Occupational Safetyand Health) (1994). In ternational Chemical Safety Card, Chloroacetic Acid, ICSC: 2035, IPCS, CEC. Online: http://www.cdc.gov/niosh/ipcsneng/neng0235.html. NOTOX (2000). Evaluation of the ability of T-7524 to induce chromosome aberrations in cultured peripheral human lymphocytes. NOTOX ProjectNumber 292062. Hertogenbosch, The Netherlands. Nubbe ME, Adams VD, and Moore WM (1995). The direct and sens itized photooxidation of hexachlorocyclopentadiene. Wat. Res. 29 , 1287-1293. Oakes KD, Sibley PK, Solomon KR, Mabury SA, and van der Kraak GJ (2004). Impact of perfluorooctanooic acid on Fathead minnow ( Pimephales promelas ) fatty acyl-CoA oxidase activity, circulating steroids, and re production in outdoor microcosms. Environ Toxicol Chem, 23 (8), 1912- 1919. Obourn JD, Frame SR, Bell RH Jr, Longnecker DS , Elliott GS, and Cook JC (1997). Mechanisms for the pancreatic oncogenic effects of the pe roxisome proliferator Wyeth-14,643. Toxicol Appl Pharmacol 145 , 425-436. OECD (2004). Environment Directorate, OECD, “ PFOS, PFAS, PFOA and Related Chemicals: Responses to the Questionnaire,” report ENV/JM 22. O’Flaherty EJ (1981). Toxicants and Drugs: Kinetics and Dynamics. John Wiley and Sons, New York. Öko-Test (2005). Test: Goretex-Jacken und Impr ägniersprays. Nicht Jacke wie Hose. Heft 5 , 158- 161. Ogata Y, Tomizawa K, and Furuta K (1983). Chemistry of Peroxides. In : S. Patai (ed) The Chemistry of Peroxides, 720.

191OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Oliaei F, Kries D, and Kessler K (2006). Investig ation of perfluorochemical (PFC) contamination in Minnesota, Phase one. Report to Minnesota Senate Environment and Natural Resources Committee. Olivero-Verbel J, Tao L, Johnson-Restrepo B, Gu ette-Fernandez J, Baldiris-Avila R, O’byrneHoyos I, and Kannan K (2005). Perfluorooctanesulfona te and related fluorochemicals in biological samples from the north coast of Colombia. Environ Poll xx, 1-6. Olsen GW, Gilliland FD, Burlew MM, Burris JM, Mandel JS, and Mandel JH (1998a). An epidemiologic investigation of reprodu ctive hormones in men with occ upational exposure to perfluorooctanoic acid. JOEM 40 , 614-622. Olsen GW, Burris JM, Burlew MM, and Mandel JH (1998b). 3M Final re port: an epidemiologic investigation of plasma cholecyst okinin, hepatic function and serum pe rfluorooctanoic acid levels in production workers. 3M Company. September 4, 1998. U.S. Environmental Protection Agency Administrative Record 226-0476. Olsen GW, Burris JM, Mandel JH, and Zobel LR (1998c ). An epidemiologic investigation of clinical chemistries, hematology and hormones in relati on to serum levels of pe rfluorooctane sulfonate in male fluorochemical production employees. 3M Company. Final Report. August 22, 1998. U.S. Environmental Protection AgencyAdministrative Record 226-0030. Olsen GW, Logan PW, Simpson CA, Hansen KJ , Burris JM, Burlew MM, Schumpert JC, and Mandel JH (1999). Fluorochemical exposure assessm ent of Decatur chemical and film plant employees. 3M Company. Final Report. August 11, 1999. U.S. Environmental Protection Agency Administrative Record 226-0950. Olsen GW, Burris JM, Burlew MM, and Mandel JH (2000). Plasma cholecystokinin and hepatic enzymes, cholesterol and lipoproteins in ammoni um perfluorooctanoate production workers. Drug Chem Tox 23 , 603-620. Olsen GW, Logan PW, Simpson CA, Burris JM, Burlew MM, Lundberg JK, and Mandel JH (2001a). Descriptive summary of serum fluorochemi cal levels among employee participants of the year 2000 Decatur fluorochemical medical surv eillance program. 3M Company. Final Report. March 19, 2001. U.S. Environmental Protecti on Agency Administrative Record 226-1030a020a. Olsen GW, Schmickler MN, Tierens JM, Logan PW, Burris JM, Burlew MM, Lundberg JK, and Mandel JH (2001b). Descriptive summary of serum fluorochemical levels among employee participants of the year 2000 Antwerp fluorochemical me dical surveillance program. 3M Company. Final Report. U.S. Environmental Protection Ag ency Administrative Record 226-1030a020b. Olsen GW, Madsen DC, Burris JM, and Mandel JH (2001c). Descriptive summary of serum fluorochemical levels among 236 building employees. 3M Company. Final Report. March 19, 2001. U.S. Environmental Protection Agency Ad ministrative Record 226-1030a020c. Olsen GW, Hansen KJ, Clemen LA, Burris JM, and Mandel JH (2001d). Id entification of Fluorochemicals in Human Tissue. 3M Company. Final Report. June 25, 2001. U.S. Environmental Protection Agency Administrative Record 226-1030a022. Olsen GW, Burlew MM, Burris JM, and Mandel JH (2001e). A cross-sectional analysis of serum perfluorooctanesulfonate (PFOS) and perfluorooctanoate (PFOA) in relation to clinical chemistry, thyroid hormone, hematology and urin alysis results from male and female employee participants of the 2000 Antwerp and Decatur fluorochemical medi cal surveillance program. 3M Company. Final Report. October 11, 2001. U.S. Environmental Prot ection Agency Administrative Record 226-1087.

192OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Olsen GW, Burlew MM, Burris JM, and Mandel JH (2001f). A longitudinal analysis of serum perfluorooctanesulfonate (PFOS) and pe rfluorooctanoate (PFOA) levels inrelation to lipid and hepatic clinical chemistry test results from male em ployee participants of the 1994/95, 1997, and 2000 fluorochemical medical surveillance program. 3M Company. Final Report. October 11, 2001. U.S. Environmental Protection AgencyAdministrative Record 226-1088. Olsen GW, Burlew MM, Hocking BB, Skratt JC, Burris JM, and Mandel JH (2001g). An epidemiologic analysis of episodes of care of 3M Decatur ch emical and film plant employees, 1993-1998. 3M Company. Final Report. May 18, 2001. U.S. Environmental Prot ection Agency Administrative Record 226-1030a021. Olsen GW, Burris JM, Lundberg JK , Hansen K, Mandel JH, and Z obel LR (2002a). Identification of fluorochemicals in human sera. I. American Red Cross adult blood donors. 3M Company. Final Report. February 25, 2002. U.S. Environmental Pr otection Agency Administrative Record 226- 1083. Olsen GW, Burris JM, Lundberg JK; Hansen KJ, Mandel JH, and Zobel LR (2002b). Identification of fluorochemicals in human sera. II. Elderly participants of the Adult Changes in Thought Study, Seattle, Washington. 3M Company. Final Report. February 25, 2002. U.S. Environmental Protection Agency Administrative Record 226-1084. Olsen GW, Burris JM, Lundberg JK, Hansen KJ, Mandel JH, and Zobel LR (2002c). Identification of fluorochemicals in human sera. III. Pediatric part icipants in a Group A Stre ptococci clinical trial investigation. 3M Company. Final Report. Febr uary 25, 2002. U.S. Environmental Protection Agency Administrative Record 226-1085. Olsen GW and Mandel JH (2003a). Descriptive anal ysis of serum fluoroche mical concentrations from Cottage Grove employee participants ofthe 2002 Medical Surveillance Program. 3M Company. Final Report. U.S. Environmental Protec tion Agency Administrative Record 226-1352. Olsen GW and Mandel JH (2003b). Descriptive anal ysis of serum fluoroche mical concentrations from Decatur employee participants of the 2002 Medical Surveillance Program. 3M Company. Final Report. U.S. Environmental Protecti on Agency Administrative Record 226-1353. Olsen GW, Schmickler M-N, Pays S, Young KA, and Mandel JH (2003c). Descriptive analysis of serum perfluorooctanoate (PFOA) and perfluorooctan esulfonate (PFOS) concentrations of Antwerp employee participants from the 2003 fluorochemi cal medical surveillance program. Final report. August 14, 2003. U.S. Environmental Protectio n Agency Administrative Record 226-1719. Olsen GW, Church TR, Miller JP, Burris JM, Ha nsen KJ, Lundberg JK, Armitage JB, Herron RM, Medhdizadehkashi Z, Nobiletti JB, O’Neill EM, Mandel JH, and Zobel LR (2003d). Perfluorooctanesulfonate and other fluoroche micals in the serum of Americ an Red Cross adult blood donors. Environ Health Perspect 111 , 1892-1901. Olsen GW, Burris JM, Burlew MM, and Mandel JH (2003e). Epidemiologic assessment of worker serum perfluorooctanesulfonate (PFOS) and perfl uorooctanoate (PFOA) concentrations and medical surveillance examinations. JOEM 45 , 260-270. Olsen GW, Butenhoff JL, and Mandel JH (2003f). Assessment of lipid, hepatic and thyroid function in relation to an occupational biologic limit valu e for perfluorooctanoate. 3M Company. Final Report. June 9, 2003. U.S. Environmental Prot ection Agency Administrative Record 226-1351. Olsen GW, Church TR, Larson EB, van Belle G, Lundberg JK, Hansen KJ, Burris JM, Mandel JH, and Zobel LR (2004a). Serum concentrations of pe rfluorooctanesulfonate and other fluorochemicals in an elderly population from Seattle, Washington. Chemosphere 54 , 1599-1611.

193OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Olsen GW, Church TR, Hansen KJ, Butenhoff JL , Mandel JH, and Zobel LR (2004b). Quantitative evaluation of perfluorooctanesulfona te (PFOS) and other FC’s in th e serum of children. J Children’s Health 2 , 1-24. Olsen GW, Burlew MM, Marshall JC, Burris JM, and Mandel JH (2004c). An alysis of episodes of care in a perfluorooctanesulfonyl fl uoride production facility. JOEM 46 , 837-846. Olsen GW, Ehresman D, Froehlich J, Burris J, a nd Butenhoff J (2005a). Eval uation of the half-life (t 1/2 ) of elimination of perfluorooc tanesulfonate (PFOS), perfluorohe xanesulfonate (PFHS) and perfluorooctanoate (PFOA) from human serum. In : Fluoros. Toronto, Canada. 18-20 August 2005. Toronto, Canada.University of Toronto TOXC017. Available:http://www.chem.utoronto.ca/symposium/fluoros/ abstractbook.htm. Olsen GW, Huang H, Helzlsouer KJ, Hansen KJ, Butenhoff JL, and Mandel JH (2005b). Historical comparison of perfluorooctanesulfonate, perfluor ooctanoate, and other fluorochemicals in human blood. Environ Health Perspect 113 , 539-545. O'Malley KD and Ebbins KL (1981). Repeat applic ation 28 day percutaneous absorption study with T-2618CoC in albino rabbits. Riker Laboratories, St. Paul, MN. U.S. E nvironmental Protection Agency Administrative Record 226-0446. Ophaug RH and Singer L (1980). Metabolic handling of perfluorooctanoic aci d in rats. Proc Soc Exp Biol Med 163 , 19-23. Pace Analytical (1987). Ready Biodegradation of FC-126(BOD/COD). 3M Company Lab Request No. E1282. Minneapolis, MN. U.S. Environmental Protection AgencyAdministrative Record 226- 0495 (May 29). Pace Analytical (2001). The 18- Day Aerobic Biodegradation Study of PerfluorooctanesulfonylBased Chemistries. 3M Compan y Request, Contract Analytical Project ID: CA097, Minneapolis, MN. U.S. Environmental Protection Agency Administrative Record 226- 1030a038 (February 23). Palazzolo MJ (1993). Thirteen-week dietary toxi city study with T-5180, ammonium perfluorooctanoate (CAS No. 3825-26-1) in male rats. Final Report. Laborat ory Project Identification HWI 6329-100. Hazleton Wisconsin, Inc. U.S. Environmen tal Protection Agency Administrative Record 226-0449. Palmer CN, Hsu MH, Griffin KJ, Raucy JL, and J ohnson EF (1998). Peroxisome proliferator acitivated receptor-alpha expression in human liver. Mol Pharmacol 53 , 14-22. Panaretakis T, Shabalina IG, Grander D, Shos han MC, and DePierre JW (2001). Reactive oxygen species and mitochondria mediate the induction of apoptos is in human hepatoma hepG2 cells by the rodent peroxisome proliferator and hepatocarci nogen, perfluorooctanoic acid. Toxicol Appl Pharmacol 173 , 56-64. Pastoor TP, Lee KP, Perri MA, and Gillies PJ ( 1987). Biochemical and morphological studies of ammonium perfluorooctanate-induced hepatomegaly and peroxisome proliferation. Exp Mol Pathol 47 , 98-109. Perkins R, Butenhoff J, Kennedy G, and Palazzolo M (2004). 13-Week dietary toxicity study of ammonium perfluorooctanoate (APFO ) in male rats. Drug Chem Tox 27 , 361-378. Plastics Europe (2005) Guide foro the Sa fe Handling of Fluoropolymer Resins, 3 rdedition, June. Prentice DE and Meikle AW (1995). A review of drug-induced Leydig cell hyperplasia and neoplasia in the rat and some comparis ons with man. Human Exper Toxicol 14 , 562-572.

194OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Powley CR, Michalczyk MJ, Kaiser MA, and B uxton LW (2005). Determination of perfluorooctanoic acid (PFOA) extractable from the surf ace of commercial cookware under simulated cooking conditions by LC/MS/MS. Analyst 130 , 1299-1302. Powley CR, Mawn M, Michalczyk MJ, Kaiser MA, Huang H, Buck R, and Korzeniowski S (2005b) Analytical methods for th e determination of ex tractable perfluorooc tanoic acid (PFOA) from consumer articles. Poster pr esentation FLUOROS meeting, Toronto Prevedouros K, Cousins I, Buck RC, and Korzenio wski SH (2006). Sources, Fate and Transport of Perfluorocarboxylates. E nviron. Sci. Technol 40 , 32-44. Ratter BA, McGowan PC, Golden NH, Hatfield JS, Toschik PC, Lukei RF, Hale RC, SchmitzAfonso I and Rice CP (2004). Contaminant Expos ure and Reproductive Success of Ospreys (Pandion Heliaetus) Nesting in Chesapeake Bay Regions of Concern. Arch Env Con Tox 47 , 126-140. Reagen WK, Lindstrom KR, Thompson KL, and Flah erty JM (2004). Analyt ical techniques and method validation for the measurement of sel ected semivolatile and nonvolatile organofluorochemicals in air. J Occup Environ Hyg 1 , 559-569. Rhoads KR, Shiang Fu Q, Higgins CP, Luthy RG, Criddle CS (2005) . Microbial transformation of N-ethyl-N-(2-hydroxyethyl)perfluorooc tanesulfonamide. Poster presente d at Fluoros: International Symposium on Fluorinated Alkyl Organic in the Environmen t, August 18-20, 2005. Toronto, Canada. Ring KL, Kalin T, and Kishi A (2002). Fluorop olymers. CEH Marketing Research Report. Risha KJ, Flaherty J, Willie R, Buck W, Morandi F, and Isemura T (2005). Method for trace level analysis of C8, C9, C10, C11, and C13 perfluorocarbon carboxylic acids in water. Anal Chem 77 , 1503-1508. Rostkowski P, Taniyasu S, Yamashita N, So MK, Lam PKS, Falandysz J, Lee KT, Seong J, Im KSH, and Giesy JP (2005). Perfluorinated compounds in streams of the Shihwa industrial zone and Lake Shihwa, South Korea. Organohalogen Compounds 67 , 767-770 Rusch G (1979). An acute inhalati on study of T-2305 CoC in the ra t. Bio/dynamics, Inc., Study No. 78-7184, May 3, 1979. U.S. Environmental Protecti on Agency Administrative Record 226-0417. Sadhu D (2002). CHO/HGPRT forward mutation assay – ISO (T6.889.7). Toxicon Corporation, Bedford, MA. Report No. 01-7019-G1, March 28, 2002. U.S. Environmental Protection Agency Administrative Record 226-1101. Saito N, Harada K, Inoue K, Sasaki K, Yosh inaga T, and Koizumi A (2004). Perfluorooctanoate and perfluorooctane sulfonate concentrations in surface water in Japan. J Occup Health 46 , 49-59. Saito, N, Harada K, Inoue K , Sasaki K , Yoshin aga T (2004b). Perfluorooctanoate and Perfluorooctane Sulfonate Concentrations in Surface Water in Japan, Journal of Occupational Health, 64 , 177. Sanderson H, Boudreau TM, Mabury SA, Cheong W-J, and Solomon KR (2002). Ecological impact and environmental fate of perfluorooctanic sulfonate on the zooplankton community in indoor microcosms. Environ Toxicol Chem 21 (7), 1490-1496. Sanderson H, Boudreau TM, Mabury SA, and Solo mon KR (2003). Impact of perfluorooctanoic acid on the structure of the z ooplankton community in indoor microcosms. Aquat Toxicol 62 , 227- 234.

195OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Sanderson H, Boudreau TM, Mabury SA, and Solom on KR (2004). Effects ofperfluorooctane sulfonate and perfluorooctanoic acid on the zooplankton community. Ecotox Environ Safety 58 , 68-76. Santoro MA (2003) submission of monitoring data pursuant to the 3M LOI dated March 13, 2003 and APFO users LOI dated March 14, 2003. U.S. E nvironmental Protection Agency Administrative Record 226-1482. August. Schrap SM, de Voogt P, van Leeuwen SP, an d Pijnenburg AMCM (2004). Perfluorinated compounds in the Dutch aquatic environment. Society of Environmental Toxicology and Chemistry Europe Annual Meeting, Prague. Poster MOPO15/008. Schroeder HF (2003). Determinati on of fluorinated surfactants a nd their metabolites in sewage sludge samples by LC with MS and tandem MS afte r pressurised liquid extr action and separation on fluorine-modified reversed-pha se sorbents. J Chromatogr A 1020 , 131-151. Schultz MM, Barofsky DF, Field JA (2003). Fluorinated Alkyl Su rfactants, Environmental Engineering Science 20 , 487-501. Schultz MM, Barofsky, DF, Field JA, (2006). Quantita tive determination of fluorinated alkyl substances by large-volume-injection liquid chromatography tandem mass spectrometry – characterization of municipal wastewaters. Environ Sci Technol 40 , 289-295.
Scott BF, Spencer C, Mabury SA, and Muir DCG (2006a). Poly and Perfluorinated Carboxylates in North American Precipitation. Environ Sci Tech web published 10/25/2006.Scott BF, Moody CA, Spencer C, Small JM, Muir DCG and Mabury SA (2006b). Analysis for Perfluorocarboxylic Acids/Anions in Surface Waters and Preci pitation Using GC-MS and Analysis of PFOA from Large-Volume Samples. Environ Sci Tech 40, 6405-6410. Scrano L, Bufo SA, Perucci P, Meallier P, a nd Mansour M (1999). Photol ysis and hydrolysis of rimsulfuron. Pestic. Sci. 55 , 955-961. Shinoda K, Hato M, and Hayashi T (1972). The physic ochensical properties of aqueous solutions of fluorinated surlitctants. J Phys Chem 76 (6), 909-914. Shoeib M, Harner T, and Vlahos P (2006). Perfluor inated Chemicals in the Arctic Atmosphere. Environ Sci Tech web published 11/17/2006. Shoeib, M., Harner, T., Wilford, B. H., Jones, K. C., and Zhu, J. (2005). Perfluorinated Sulfonamides in Indoor and Outdoor Air and Indoor Dus t: Occurrence, Partitioni ng, and Human Exposure. Environmental Science and Technology 39 , 6599-6606. Sibinski LJ (1987). Final report of a two year or al (diet) toxic ity and carcinogeni city study of fluorochemical FC-143 (perfluorooctanane amm onium carboxylate) in rats. Vol.1-4, 3M Company/RIKER Exp. No.0281CR0012; 8EHQ-1087-0394, October 16, 1987. Simister EA, Lee EM, Lu JR, Thomas RK, Otte will RH, Rennie AR, and Penfold J (1992). Adsorption of ammonium perfluoroocanoate and ammonium decanoate at the air-solution interface. J Chem Soc, Faraday Trans 88 , 3033-3041. Simons JH (1949). Electrochemical Process for the Production of Fluorocarbons. J. Electrochem. Soc

. 95 , 47-59. Sinclair E, Taniyasu S, Yamashita N, and Kannan K (2004). Perfluor ooctanoic acid and perfluorooctane sulfonate in Michigan and New York waters. Or ganohalogen Compounds 66 , 4069- 4073.

196OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Sinclair E and Kannan K (2006a). Mass Loading of Perfluoroalkyl Surfactants in Wastewater Treatment Plants, Environ Sci Technol 40 , 1408-1414. Sinclair E, Mayack D, Roblec K, Yamashita N, and Kannan K, (2006). O ccurrence of Perfluoroalkyl Surfactants in Water, Fish and Birds from New York State. Arch Environ Contam Toxicol 50 , 398-410. Skutlarek D, Exner M and Färber H (2006). Perfl uorinated Surfactants in Surface and Drinking Waters. Enviro Sci Pollut Res 13(5): 299-307. Smithwick M, Mabury S, Solomon K, Sonne C, Martin JW, Born EW, Dietz R, Derocher AE, Letcher RJ, Evans TJ, Gabrielsen GW, Nagy J, S tirling I, Taylor MK, and Muir DCG (2005a). Circumpolar study of perfluoroalkyl contaminants in polar bears (Ursus martimus). Environ Sci Technol ASAP Web Release Date: 22-Jun-2005; (Artic le) http://pubs3.acs.org/ acs/journals/doilookup/in doi=10.1021/es048309w. Smithwick M, Muir DCG, Mabury SA, Solomon KR, Martin JW, Sonne C, Born EW, Letcher J, and Dietz R (2005b). Perfluoroalkyl contaminants in liver tissue from east Greenland polar bears (Ursus maritimus). Environ Toxicol Chem 24 , 981-986. Smithwick M, Norstrom RJ, Mabury SA, Solomon K, Evans TJ, Stirling I, Taylor MK, and Muir DCG (2006). Temperoal trends of perfluoroalkyl contaminants in polar bear (Ursus maritimus) from two locations in the North Americ an Arctic, 1972-2002). Environ Sci Technol 40 , 1139-1143. SNF (2006). Fluorinated pollutants in all-weather clothing. Swedis h Society for Nature Conservation 43 pages. So MK, Taniyasu S, Yamashita N, Giesy JP, Zh eng J, Fang Z, Im SH, and Lam PKS (2004). Perfluorinated compounds in coastal waters of Hong Kong, South China, and Korea. Environ Sci Technol 38 , 4056-4063. So MK, Taniyasu S, Lam PKS, Zheng GJ, Giesy JP , and Yamashita N (2005a). Perfluorinated acids in mussels and oysters from South China and Japan. Organohalogen Compounds 67 , 738-740 So MK, Taniyasu S, Lam PKS, Zeng GJ, Giesy JP, and Yamashita N (2006a). Alkaline digestion and solid phase extraction method for perfluorinated compounds in mussels and oysters from South China and Japan. Arch Environ Contam Toxicol 50 , 240-248. So, MK, Yamashita N, Taniyasu S, Jiang Q, Gies y JP, Chen K, and Lam PK (2006b). Health risks in infants associated with exposure to perfluorin ated compounds in human breast milk from Zhoushan, China. Environ Sci Technol 40, 2924-2929. Sohlenius A-K, Andersson K, and DePierre JW ( 1992). The effects of perfl uoro-octanoic acid on hepatic peroxisome proliferation and related parameters show no sex-related differences in mice. Biochem J 285 , 779-783. Sonich-Mullin C, Fielder R, Wiltse J, Baetcke K, Dempsey J, Fenner-Crisp P, Grant D, Hartley M, Knaap A, Kroese D, Mangelsdorf I, Meek E, Rice JM, and Younes M (2001). IPCS conceptual framework for evaluating a MOA for chemical carcinogenesis. Regul Toxicol Pharmacol 34 , 146- 152. Southern Research Institute (2003) . Protein binding of perfluorohe xane sulfonate, perfluorooctane sulfonate and perfluoroo ctanoate to plasma (human, rat an d monkey), and various human derived plasma protein fractions. Study ID: 9921.7. U.S. Envi ronmental Protection Agency Administrative Record 226-1354.

197OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

SPI (2002). Letter of Society of the Plas tics Industry to US EPA, April 26 th , 2002. US EPA Docket AR 226-1094. SPI (2005). The Society of the Plas tics Industry: Dispersion Processor Material Balance Project. OPPT-2003-0012-0900 through -0904. 2005/02. Staples RE, Burgess BA, and Kerns WD (1984). The em bryo-fetal toxicity and teratogenic potential of ammonium perfluorooctanoate (APFO ) in the rat. Fundam Appl Toxicol 4 , 429-440. Starkov AA and Wallace KB (2002). Structural determinants of fluorochemical-induced mitochondrial dysfunction. Toxicol Sci 66 , 244-252. Stock NL, Lau FK, Ellis DA, Martin JW, Muir DCG and Mabury SA (2004). Polyfluorinated Telomer Alcohols and Sulfonamides in the North Am erican Troposphere. Envi ron Sci Technol 38(4): 991-996. Strynar, MJ and Lindstr om, AB (2005) Perfluorinated compounds in archived house dust samples. Poster presented at Fluoros: In ternational Symposium on Fluorinated Alkyl Organic in the Environment, August 18-20, 2005. Toronto, Canada. Sykes G (1987). Two-year toxicology/carcinogenici ty study of fluorochemical FC-143 in rats. Memo from G Sykes to C Reinhardt, Hask ell Lab. Toxicol. Ind. Med., October 29, 1987. Takagi A, Sai K, Ummemura T, Hasegawa R, an d Kurokawa Y (1991). Short-term exposure to the peroxisome proliferators, perfluorooctanoic acid and perfluorodecan oic acid, causes significant increases of 8-hydroxydeoxyguanosine in liver DNA of rats. Cancer Lett 57 , 55-60. Takagi A, Sai K, Ummemura T, Hasegawa R, and Kurokawa Y (1992). Hepatomegaly is an early biomarker for hepatocarcinogenesis i nduced by peroxisome proliferat ors. J Environ Toxicol Pathol 11 , 45-149. Taniyasu S, Yamashita N, Kannan K, Horii Y, Sinc lair E, Petrick G, and Gamo T (2004). Perfluorinated carboxylates and sulfonates in open ocean wa ters of the Pacific and Atlantic Oceans. Organohalogen Compouds 66 , 4035-4040. Taniyasu S, Kannan K, So MK, Gulkowska A, Si nclair E, Okazawa T, and Yamashita N (2005). Analysis of fluorotelomer alcohol s, fluorotelomer acids, and short- and long-chain perfluorinated acids in water and biot a. J Chromatography A 1093 : 89-97. Taniyasu S, Kannan K, So MK, Gulkowska A, Si nclair E, Okazawa T, and Yamashita N (2005a). A new solid phase extraction method for short- and l ong-chain perfluorinated acids and fluorotelomers in water and biota. Organohalogen Compounds 67 , 760-763 Teerds KJ, Rommerts FG, and Dorrington JH (1990) . Immunohistochemical detection of transforming growth factor-á in Leydig cell during the deve lopment of the rat testis. Molec Cell Endocrinol 69 , R1-R6. Theobald N and Caliebe C (2005). Entwicklung und Validierung einer Methode zur Bestimmung von perfluorierten organischen Substanzen im M eerwasser, Sedimenten und Biota; Untersuchungen zum Vorkommen dieser Schadstoffe in der No rd- und Ostsee. UBA F & E-Vorhaben 202 22 213. 4. Zwischenbericht: Zeitraum: 01.07.2004-31.12.2004. Thomford PJ (2001a). 4-Week capsule toxicity study with ammonium perf luorooctanoate (APFO) in cynomolgus monkeys. Study performed by C ovance Laboratories Inc., Madison Wisconsin 53704-2592 for APME Ad-hoc APFO Toxicol ogy Working Group. Study No. Covance 6329-230,

198OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Completion Date December 18, 2001, 159 pp. U.S. Envi ronmental Protection Agency Administrative Record 226-1052a. Thomford PJ (2001b). 26-Week capsule toxicity study with ammonium perf luorooctanoate (APFO) in cynomolgus monkeys. Study performed by C ovance Laboratories Inc., Madison Wisconsin 53704-2592 for APME Ad-hoc APFO Toxicol ogy Working Group. Study No. Covance 6329-231, Completion Date December 18, 2001, 463 pp. U.S. Envi ronmental Protection Agency Administrative Record 226-1052a. Tominaga N, Kohra S, Iguchi T, and Arizono K (2004). Effects of perfluoro organic compound toxicity on nematode Caenorhabditis elegans fecundity. J Health Sci 50(5), 545-550. Tomy GT, Budakowski W, Halldorson T, Helm PA, Stern GA, Friesen K, Pe pper K, Tittlemier SA, and Fisk AT (2004). Fluorinated or ganic compounds in an eastern arctic marine food web. Environ Sci Technol 38 , 6475-6481. Tomy G, Halldorson T, and Tittlemier S (2005). Meth anolic extraction of poly- and perfluorinated alkyl acids from biota. Organohalogen Compounds 67 , 787-789 Trosko JE, Chang CC, Upham BL, and Wilson M (1998). Epigenetic toxicology as toxicantinduced changes in intracellular signaling leading to altered gap junctional intercellular communication. Toxicol Lett 102-103 , 71-78. Tseng C-L, Liu L-L, Chen C-M, and Ding W-H ( 2006) Analysis of perfl uorooctanesulfonate and related fluorochemicals in water and biological tissue samples by liquid chromatography-ion trap mass spectrometry. J Chromatography A 1105, 119-126 TRP (2002). Telomer Research Program Update - Presented to the USEPA OPPT, November 25, 2002. USEPA Administrative Record AR226-1141. 2002/11/25. TRP (2003). Telomer Research Program Letter of Intent. Letter 14-March, U.S. EPA EDocket OPPTS-2003-0012-0013. Tugwood JD, Holden PR, James NH, Prince RA, and Roberts RA (1998). A PPAR alpha cDNA cloned from guinea pig liver enc odes a protein with similar prope rties to the mouse PPAR alpha: Imprications for species differences in respons e to peroxisome proliferators. Arch Toxicol 72 , 169- 177. UBA (2004). PFOS and PFOA in human blood plas ma 1985-2004. Human section of the German Environmental Specimen Bank, unpublished study. Ubel FA, Sorenson SD, and Roach DE (1980). Health status of pl ant workers exposed to fluorochemicals--a preliminary report. Am Ind Hyg Assoc J 41 , 584-589. Upham BL, Deocampo ND, Wurl B, and Trosko JE (1998). Inhibition of gap junctional intercellular communication by perfluorinated fatty acids is dependent on the chain length of the fluorinated tail. Int J Cancer 78 , 491-495. USEPA. (1992). Atmospheric Modelling Stud ies. USEPA Administ rative Record AR226-2423. USEPA. (1993). C-8 Review. February 8, 1993.http://www.epa.gov/oppt/tsca8e/pubs/8ehq/2005/august05/8ehq-1104-15856a88050000062s/vol_2-4/226-2070.pdf

199OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

USEPA (1998). Fate, Transport and Transforma tion Test Guidelines, OPPTS 835.5270 Indirect Photolysis Screening Test; EPA712-C-98-099; Un ited States Environmental Protection Agency, U.S. Government Printing Office: Washington, DC, 1998, 1-22. US EPA (2002). Draft Hazard assessment of perf luorooctanoic acid and its salts. U.S Environmental Protection Agency, Office of Pollution, Pr evention and Toxics, Risk Assessment Division. November 4, 2002 USEPA (2002a). Memorandum from Dr. Ralph Coope r, NHEERL, to Dr. Jennifer Seed, dated October 2, 2002. USEPA (2002b). Memorandum from Dr. Elizabeth Margosches to Dr. Katherine Anitole, dated October 21, 2002. USEPA (2003). Letter from Robert A. Bilott to Ri chard H. Hefter, dated July 3, 2003. U.S. Environmental Protection Agency Admini strative Record 226-1372; 8EHQ-0703-00373. USEPA (2003b). Enforceable Consent Agreement for the Laboratory Scale Incineration Testing of Fluoropolymers. EPA Do cket OPPT-2003-0071. USEPA (2005). U.S. Environmental Protection Agen cy, PBT Profiler, “Chemicals that Can’t be Profiled: Highly Fluorinated Compounds,” www.pbtprofiler.net/chemicals.asp#Fluor. Van de Vijver KI, Hoff PT, Das K, Van Dongen W, Esmans EL, Jauniaux T, Bouquegneau J-M, Blust R, and De Coen W (2003). Perfluorinated chemicals infiltrate ocean waters: Link between exposure levels and stable isotope ratios in marine mammals. Environ Sci Technol 37 , 5545-5550. Van de Vijer KI, Hoff PT, Das K, Van Dongen W, Esmans EL, Si ebert U, Bouquegneau JM, Blust R, and De Coen WM (2004). Baseline study of perf luorochemicals in harbor porpoises (Phocoena phocoena) from Northern Europe. Mar Pollut Bull 48 , 986-1008. Van Leeuwen S, Karrman A, Zammit A, Van Bavel B, Van der Veen I, Kwadijk C, de Boer J, and Lindstrom G (2005). 1 st Worldwide Interlaboratory study on perfluorinated compounds in human and environmental matrices. Report, August 11, 2005. Abstract available in: Fluoros. Toronto, Canada. 18-20 August 2005. Toronto, Canada. To ronto, Canada:University of Toronto ANA040.Available:http://www.ch em.utoronto.ca/symposium/ fluoros/abstractbook.htm. Vanden Heuvel JP, Kuslikis BI, and Peterson RE ( 1991a). Covalent binding of perfluorinated fatty acids to proteins in the plasma, liver and testes of rats. Chem-Biol Interact 82 , 317-328. Vanden Heuvel JP, Kuslikis BI, Van Rafelghem ML, and Peterson RE (1991b). Tissue distribution, metabolism, and elimination of perfluorooctanoic acid in male and female rats. J Biochem Toxicol 6 , 83-92.Vanden Heuvel JP, Davis JW, Sommers R, and Peterson RE (1992). Renal excretion of perfluorooctanoic acid in male rats: Inhibitoryeffect of testoster one. J Biochem Toxicol 7 , 31-36. Wallington TJ, Hurley MD, Xia J, Wuebbles DJ, Sillman S, Ito A, Penner JE, Ellis DA, Martin J, Mabury SA, Nielsen OJ and Sulbaek AndersenMP (2006). Formation of C7F15COOH (PFOA) and Other Perfluorocarboxylic Ac ids during the Atmospheric Oxidat ion of 8:2 Fluorotelomer Alcohol. Environ. Sci. Technol. 40(3): 924-930. Walrath J and Burke C (1989). An investigation in to the occurrence of leukemia at Washington Works. E.I. Dupont De Nemours and Company, April 1989. U.S. Environmental Protection Agency Administrative Record 226-1308-1.

200OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Wang N, Szostek B, Folsom PW, Sulecki LM, Capka V, Buck RC, Berti WR, Gannon JT, (2005a) Aerobic Biotransformation of 14C-Labeled 8: 2 Telomer B Alcohol by Activated Sludge from a Domestic Sewage Treatment Plant. E nviron. Sci. Technol. 39(2); 531-538. Wang, N., Szostek, B., Buck, R. C., Folsom, P. W ., Sulecki, L. M., Capka, V., Berti, W. R., Gannon, J. T., (2005b) Fluorotelomer Alcohol Biodegradation-Direct Ev idence that Perfluorinated Carbon Chains Breakdown. Envir on. Sci. Technol., 39(19), 7516-7528. Washburn ST, Bingman TS, Braithwaite SK, BuckRC, Buxton LW, Clewell HJ, Haroun LA, Kester JE, Rickard RW, and Shipp AM (2005). Exposure assessment and risk characterization for perfluorooctanoate in selected consumer articles. Environ. Sci. Technol. 39 , 3904-3910 (Supplement). Wendling L (2003). Memo to Stephen L Johnson O PPT US EPA Re: Environmental, Health and safety measures relating to PFOA and its salts. EPA-HQ-OPPT-2003-0012-0007. March 13, 2003. Wendling L (2004). Memorandum of understanding Between the US Environmental ProtectionAgency and 3M company and Dyneon LLC for a perfluorooctanoic acid (PFOA) site-related environmental assessment program. October 12, 2004. EPA-HQ-OPPT-2004-0112-0002. Welter AN (1979). Technical report summary - fina l comprehensive report: FC-143. 3M Company. U.S. Environmental Protection Agency Administrative Record 226-528. Will R, Kälin T and Kishi A (2005). Fluoropolymer s CEH Marketing Research Report SRI International, Menlo Park, CA. Yager JD Jr and Yager R (1980). Or al contraceptive steorids as pr omoters of hepatocarcinogenesis in female Sprague-Dawley rats. Cancer Res 40 , 3680-3685. Yamada T, Taylor PH, Buck RC, Kaiser MA, a nd Giraud RJ (2005). Thermal degradation of fluorotelomer treated articles and related materials. Chemosphere 61, 974-984. Yamashita N, Kannan K, Taniyasu S, Horii Y, Okazawa T, Petrick G, and Gamo T (2004a). Analysis of perfluorinated acids at parts-per-quadrillion levels in seawater using liquid chromatographytandem mass spectrometry. Environ Sci Technol 38 , 5522-5528. Yamashita N, Kannan K, Taniyasu S, Horii Y, Hanari N, Okazawa T, and Petrick G (2004b). Environmental contamination by perflu orinated carboxylates and sulfona tes following the use of firefighting foam in Tomakomai, Japan. Organohalogen compounds 66 , 4063-4068. Yamashita N, Kannan K, Taniyasu S, Horii Y, Pe trick G, and Gamo T (2005). A global survey of perfluorinated acids in oceans. Mar Pollut Bull 51, 658-668 Yamazaki Y, Harada K, Morikawa A, Yoshinag a T, and Koizumi A ( 2005). Prolongation of K+- induced backward swimming of Paramecium caudatum by PFOS and PFOA. The Toxicologist 84 (S-1), 423 (Abstract 2067). Yang Q, Xie Y, and Depierre W (2000). Effects of peroxisome proliferators in the thymus and spleen of mice. Clin Exp Immunol 122 , 219-226. Yang Q, Xie Y, Ericksson AM, Nelson BD, and De Pierre JW (2001). Furthe r evidence for the involvement of inhibition of cell pr oliferation and devel opment in thymic and splenic atrophy induced by the peroxisome proliferator perfluoroc tanoic acid in mice. Biochem Pharmacol 62 , 1133-1140. Yang Q, Xie Y, Alexson SHE, Nelson BD, and De Pierre JW (2002a). Involvement of the peroxisome proliferator-activated receptor alpha in th e immunomodilation causedby peroxisome proliferators in mice. Biochem Pharmacol 63 ,1893-1900.

201OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Yang Q, Abedi-Valugerdi M, Xie Y, Zhao X, Mo ller G, Nelson BD, and DePierre JW (2002b). Potent suppression of the adaptive immune respons e in mice upon dietary exposure to the potent peroxisome proliferator, perfluorooctano ic acid. Internat Immunopharmacol 2 , 389-397. Ylinen M, Hanhijarvi H, Jaakonaho I, and Peura P (1989). Stimulation by estradiol of the urinary excretion of perfluorooctanoic acid in the male rat. Pharmacol Toxicol 65 , 274-277. Ylinen M, Kojo A, Hanhijarvi H, and Peura P (1990). Dispositionof perfl uorooctanoic acid in the rat after single and subchronic administration. Bull Environ Contam Toxicol 44, 46-53. York RG (2002). Oral (gavage) two-generation (one litter per ge neration) reproduction study of ammonium perfluorooctanoic (APFO) in rats. Argus Research Laborat ories, Inc. Protocol Number: 418-020, Sponsor Study Number : T-6889.6, March 26, 2002. U.S. E nvironmental Protection Agency Administrative Record 226-1092. Young, C.J., Muir, D.C.G., Koerner, R.M., an d Mabury, S.A. 2005. Atmospheric Flux of Perfluorinated Acids into the High Arctic. Poster pres entation at the Internati onal Fluoros Symposium. Toronto, Ontario, Canada. August 19-20, 2005. Zipfel, RJ (1987). C-8 Reduc tion & Control Programs.http://www.epa.gov/oppt/tsca8e/pubs/8ehq/2005/august05/8ehq-1104-15856a88050000062s/vol_8-9/226-2393.pdf

202OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

ANNEX

Analytical Methods

Table 50 Analysis of air samples Sample Sample preparation Analysis Limit of detectionand/ or quantification// Recovery Reference

Air Adsorption on Amberlite XAD2 cartridges; elution with methanol or ethyl acetate GC/MS LC/MS/MS 19 F NMR NR Ellis et al. 2004

Air Collection of air samples viaOSHA versatile sampler consisting of glass or quartz fiber filter followed by a polystyreneresin sorbent; samples were collected at a flowrate of 1 L/min; collection times 8 h; sample divided into three fractions (plastic retaining ring and glassor quartz fiber filter; first section of resin beads and firstpolyurethane foam; second section of resin beads and the back PUF filter); a. fractions were agitated with C 10 solution and methanol; extracts were transferred to other vials and analysed b. fraction were placed in glass tubes and methanol added; the mixture was shaken for 60 min; an aliquot was filtered and transferred to a amber glass vial; 1,2- 13 C-perfluorooctanoic acid as internal standard Electrospray LC/MS or LC/MS/MS LOQ ; 0.474 μg/m3// 88-95.7% Kaiser et al., 2005a

Air (particulateand vapor samples from workplace air) Air collected on OVS (OSHAversatile sampler) tubes containing glass fiber filter, XADand polyurethane foam; extraction with acetone LC/MS LOQ: 8 μg/m 3 LOD: 2.4 μg/m 3 //
94% (C 8 F 15 OONH 4 ) Reagen et al. 2004

Dust Samples were extracted with methanol in a centrifuge tube using ultrasonic agitation; after centri-fugation the supernatant was filtered; for analysis, the internal standard in methanol was added to an aliquot of the filtrate and then analyzed LC/MS/MS LOD: 0.5 ng/ml LOQ: 50 ng/g//
73% Moriwaki et al. 2003

203OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 51 Analysis of water samples Sample Sample preparation Analysis Limit of detectionand/ or quantification// Recovery Reference

Surface water, drinking water Water sample was extracted via SPE at pH 7-8; after drying the cartridge was extracted with aceton/acetonitrile; the extracts were dried; the residues were dissolved in methanol/water/ ammonium acetate HPLC/MS/ MS LOD: 2 ng/l// 50-95% Skutlarek, Exner, and Färber 2006

Surface water Water samples were extracted using a weak anion exchange cartridge; the analytes were eluted with methanol and 0.1% ammonium methanol HPLC/MS/ MS LOQ : 0.1 ng/l// NR Falandysz et al., 2005

Surface water Water samples were mixed with an equal volume of methanol containing internal standards ( 13 C-PFOA and 13 C 2 -PFDA); the mixture was filtered HPLC/MS/ MS LOD: 1-4 ng/l// NR Furdui et al., 2005

Surface water Extraction of water samples was carried out using OasisHLB and Oasis WAX cartridges; cartridges were preconditioned with methanol and water (HLB) and NH 4 OH in methanol, methanol and water (WAX); samples were passed through the cartridges at a rate of 1 drop/sec; cartridges were then washed and analytes eluted HPLC/MS/ MS LOD: 0.01-1 ng/l// >80% (HLB) 85-107% (WAX) Taniyasu et al., 2005a

River water Samples were thoroughly mixed; 40 ml of the sample was transferred to a polypropylene centrifuge tube; extraction wasconducted using SPE cartridges; analytes were eluted with 100% methanol HPLC/MS/ MS LOD: 25 ng/l LOQ : 25–50 ng/l//
112–114% Hansen et al. 2002

Surface water Filtration of the water sample; the water was pushed through a solid phase extraction cartridge; elution of the cartridges with methanol HPLC/MS LOD: 6 ng/l LOQ: 13 ng/l//
102–113% Boulanger et al. 2004

204OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 51 Analysis of water samples Sample Sample preparation Analysis Limit of detectionand/ or quantification// Recovery Reference

Surface water Water samples were spiked with perfluorooctanesulfonateas inter-nal standard; the samples were extracted using Oasis HLB cartridges; elution was carried out with methanol; the eluate was reduced to 0.2–0.5ml under a gentle stream of nitrogen HPLC/MSMS LOQ: 0.02 ng/l//
160–190% So et al. 2004

water Enrichment on SPE cartridge; elution with methanol; solution concentrated under argon HPLC/cond uctometricdetection/MS andGC/MS LOD: 200 μg/l//
100% Hori et al. 2004a, b

Water (river,coastal sea water, tap water) Samples filtered through glassand membrane filter; enrichment on cartridges; elution with methanol; concentrated under stream of nitrogen HPLC/MS LOD: 0.06 ng/l LOQ: 0.1 ng/l//
92–99% Saito et al. 2004

Water (well, stream, spring, tap, Omnisolve, and type I water) Enrichment on SPE cartridges; elution with 100% methanol LC/MS/MS LOQ: 25 ng/l//
70–130% Risha et al. 2005

Surface Water Samples were extracted using SPE cartridges; the analytes were eluted with methanol LC/MS/MS LOQ: 9 ng/l//
80–109% Moody et al. 2002 Moody et al. 2001

Lake water Thawed samples were allowed to settle; aliquots were decanted into polypropylene bottles and spiked with PFBS recovery standard; extraction was carriedout via SPE cartridges; cartridges were eluted with methanol and the eluent concentrated HPLCMS/MS LOD: ca. 8 ng/l LOQ: 25 ng/l//
NR Sinclair et al. 2004

205OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 51 Analysis of water samples Sample Sample preparation Analysis Limit of detection and/ or quantification// Recovery Reference

Water samples(river water, effluent of industrial/ municipal WWTP; and coastal water) Water samples were adjusted to pH 2-3; preconcentrated with RP-C 18 SPE cartridges; the cartridge was washed with deionised water; elution was carried out using tert-butyl ether/methanol ((1:1), MTBE/methanol (1:4), and methanol alone; extracts were evaporated to less than 100 μlby a stream of nitrogen; addition of deionised water resulting in a final volume of 1 ml; addition of internal standard (PFDoA) HPLC/MS LOD: 0.8 ng/l LOQ: 2 ng/l// 52-103% Tseng et al., 2006

Water (effluent of WWTP) Samples were allowed to settle;an aliquot of 200 ml was decanted into a polypropylenebottle; addition of internal standard (PFBS and 13 C-PFOA); the mixture was passed through cartridge; after this extraction the cartridge was washed with 20% aqueous methanol and Milli-Q water; after drying ofthe cartridge the target compounds were eluted withmethanol; the extract was concentrated and filtered HPLC/MS/ MS LOQ : 2.5 ng/l//70-130% (for different fluorochemicals e.g. PFOA) Sinclair & Kannan 2006

Water samples Samples were acidified with formic acid and filtered; a. the filtered particles were spiked with recovery standard and extracted via ASE (see sediment and sludge samples)
b. ammonium acetate and internal standard were added to the water phase; the water sampleswere extracted using SPE cartridges; the cartridges were then rinsed with methanol/water(40:60) and eluted with methanol; after concentrating the methanol extract, recovery standard and aqueous buffer was added HPLC/MS LOD: 0.06 ng/l LOQ: 3.0 ng/l//
NR Berger et al. 2004
Kallenborn et al. 2004

206OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 51 Analysis of water samples Sample Sample preparation Analysis Limit of detection and/ or quantification// Recovery Reference

Water (effluentof urban/industrial WWTP and seawater) Ion-pair SPME GC/MS LOD: 0.1 ug/l LOQ: 0.34 ug/l//
73–90% Alzaga & Bayona 2004

STP effluent and biosolids a. Biosolids were air dried andreconstituted with distilled water; after addition of internal standard, sodium carbonate andtetrabutylammonium hydrogensulfate the samples were homogenized; slurries were extracted with MTBE; samples were evaporated to dryness and reconstituted in methanol; the final extract was filtered b. effluent samples were dilutedwith methanol containing internal standards; the solution was filtered HPLC/MS/ MS LOD: 1 ng/l and 0.1 ng/g dw// NR Crozier et al., 2005

wastewater Water sample was treatedwith sodium thiosulfate solution; 40 ml sample was transferred to a C 18 -SPE cartridge; cartridgewas washed with 40% methanol/water; analytes were eluted with 100% methanol. HPLC/MS/ MS LOQ : 50 ng/l// 98-100% Exygen 2006

Plant effluent Samples filtered through glassfiber filter; SPE on C18 column; column dried in N2- stream; elution with methanol; extract reduced on a rotationvacuum concentrator; redissolved in methanol; addition of internal standard (1H,1H, 10H,10H-hexadecafluoro-1,10- decandiol) HPLC/MS LOD: 25 ng/l LOQ: 50 ng/l//
103% Hohenblum et al. 2003

Seawater, open ocean water Samples were spiked with perfluorooctanesulfonate as internal standard; water sampleswere extracted using SPE cartridges; elution was conducted with methanol; the solvent was removed by a gentle stream of nitrogen; particles in the finalsolution were removed by filtration HPLC/MSMS LOD: 5.2 pg/l LOQ: 1–100 ng/l//
147% Yamashita et al. 2004a Taniyasu et al. 2004

207OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 51 Analysis of water samples Sample Sample preparation Analysis Limit of detection and/ or quantification// Recovery Reference

Sea water Water samples collected in 5 m water depth; the water phase was acidified with hydrochloricacid to pH 2–3; addition of internal standard (d 6 -diuron and d 3 -mecoprop); the aqueous solution was pumped through acartridge filled with Chromabond or Cromafix HR-P resin; the cartridges were washed with pure water and dried under N2- stream; elution with methanol or methanol ammonium acetate,and acetic acid in reversed direction; the extract was concentrated and then analyzed HPLC/MS/ MS LOD: 0.05–0.5 ng/l//
97–106% Caliebe et al. 2004
Theobald and Caliebe, 2005

Sea water Samples were extracted using HLB and WAX cartridges; the cartridges were washed and the target analytes eluted HPLC/MSMS ca. 3 ng/l//
>80% Taniyasu et al., 2005

Water Samples were extracted by C 18 - SPE; cartridges were eluted with methanol; clean up via adsorption chromatography HPLC/MS LOD: 80 ng/l//
NR De Voogt and van Roon 2005

Surface water Sample was transferred to a clean flask, residual chlorine was reduced by adding sodium thiosulfate; 13 C-PFOA was added as internal standard; the samples were shaken andpassed through a cartridge; cartridge was eluted with methanol; the extract was concentrated to 1 ml and then filtered HPLC/MS/ MS LOQ : 900 ng/l//
91-114% Kurunthachalam et al., 2005

Sea water Samples were spiked with perfluorooctanesulfonate as internal standard; water samples were extracted using HLB AND WAX cartridges; elution was conducted with methanol; thesolvent was removed by a gentle stream of nitrogen; particlesin the final solution were removed by filtration HPLC/MSMS 0.08 ng/l
>90% Taniyasu et al., 2005

208OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 52 Analysis of sediment and sludge samples Sample Sample preparation Analysis Limit of detectionand/ or quantification// Recovery Reference

Sludge Air-dried sludge was spiked with PFBS and 13 C-PFOA as internal standards; the sludge was sonicated at 60°C for 20 min in 7.5 ml acetic acid; the supernatant was removed by centrifugation; the residue was resuspended in methanol/1% acetic acid and was sonicated at 60°C for 20 min; the extracts were combined; the extraction procedure was repeated threetimes; after addition of 1% acetic acid the extract was passed through cartridge; the cartridge was washed with 20% methanol and dried completely under vacuum; target compounds were eluted with methanol; after concentrating to 1 ml the extract was filtered HPLC/MS/ MS LOQ: 10 ng/g dw//70-130% (for different fluorochemicals, e.g. PFOA) Sinclair & Kannan, 2006

Sediment and sludge samples Homogenized, air dried samples were washed with acetic acid and vortexed; extraction of the samples via a mixture of methanol and 1% acetic acid (90:10); the wash and extraction was carried out in a heated (60°C) sonication bath; mixture was centrifuged and the aceticacid solution separated; the extracts were passed through C 18SPE cartridges; the fluorochemi-cals were eluted from the cartridges using methanol; after concentrating in a gentle stream of nitrogen 0.01% aqueous ammonium hydroxide solution was added; finally internal standard was added HPLC/MS/ MS LOD: 0.011 ng/g dw (sediment)//
88% LOD: 1 ng/g dw (sludge)//
71% Higgins et al., 2005

209OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 52 Analysis of sediment and sludge samples Sample Sample preparation Analysis Limit of detectionand/ or quantification// Recovery Reference

Sediment and sludge samples Samples were spiked with 7-Hperfluoroheptanoic acid as internal standard and extractedusing acclerated solvent extraction (ASE); the extraction was performed three times with methanol; then the extracts were concentrated; recovery standard (3,5-bis(trifluoromethyl) phenylacetic acid) and aqueous ammonium acetate were added; thesolution was treated in an ultrasonic bath and filtered HPLC/MS LOD: 6 ng/g ww LOQ: 200 ng/g ww//
NR Berger et al. 2004
Kallenborn et al. 2004

Sewage sludge Sludges were extracted via pressurized liquid extractionwith ethylacetatedimethylformamide and methanol-phosphoric acid as extraction solvents; the extracts were reduced in volume; no further clean-up step was necessary Flow injection analysis combined with MS
LC/MS LOD: 10 mg/kg dw//
NR Schröder 2003

Sediment and soil Samples were homogenized; addition of internal standard, tetrapropylammonium acid (pH 10), and Na 2 CO 3 /NaHCO 3buffer; tubes were closed; addition of MTBE and Cu powder;mixture was shaken and sonicated; centrifugation; MTBE extract was separated; solvent evaporated; after addition ofmethanol the mixture was filtered HPLC/MS LOD: 0.4 ng/g dw//
NR De Voogt and van Roon 2005

Sediment To the samples surrogate standard, sodium carbonate, and tetrabutylammonium hydrogen sulfate was added ; the sediment samples were extracted withMTBE; MTBE layer was separated using a centrifuge; the extracts were reduced in volume and 1,1,1,3,3,3-hexafluoro-2- propanol added; the extractswere then filtered and evaporated to dryness; extracts were reconstituted in methanol HPLC/MS LOQ/LOD: NR//
50-80% Lucaciu et al., 2005

210OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 53 Analysis of biota samples Sample Sample preparation Analysis Limit of detectionand/ or quantification// Recovery Reference

Biota (polar cod and glaucous gill liver) Liver was homogenized:
a. addition of internal standard (7-H-PFHpA), methanol/water and ammonium acetate; samplewas then vortexed and sonicated; the mixture was filtered and recovery standard was added
b. . addition of internal standard (7-H-PFHpA), water, buffer andtetra –n-butylammonium hydrogen sulfate; pH adjusted to 10; extraction performed using tert-butyl methyl ether; extracts were evaporated to dryness;residue suspended in methanol/water and recovery standardwas added; the extract was filtered HPLC/MS LOD: 1.28 ng/g ww//
83-84 % Berger and Haukas, 2005

Biota (e.g. muscle and liver tissue) Tissue samples were agitated with KOH in methanol and the mixture shaken for 16 h; theextract was diluted with distilled water; the extract waspassed through a WAX cartridge; the cartridge was washed and the analytes eluted HPLC/MS/ MS LOD: 0.01-0.08 ng/g ww//
NR Taniyasu et al., 2005a

Biota (narwhal and burbot liver) Homogenized samples were;extracted using methanol, vortexing the mixture and centrifuging; the extraction was repeated 2 more times; the supernatant methanol phases were separated; the methanol phases were combined and reduced involume; the residue was transferred to a microcentrifuge vial and ultracentrifuged; the methanol extract was removedand transferred to a HPLC injection vial; the mixture was spiked with internal standards;for comparison a ion-pair extraction was performed withhomogenized burbot liver according to Hansen et al., 2001 HPLC/MS/ MS LOD/LOQ: NR//
>85% Tomy, Halldorson, and Tittlemier, 2005

211OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 53 Analysis of biota samples Sample Sample preparation Analysis Limit of detectionand/ or quantification// Recovery Reference

Biota (organs and tissue of harbor seals) Tissue was homogenized on ice and in the presence of Milli-Q water; homogenate, internal standard, tetrabutylammonium hydrogen sulfate (pH 10), and carbonate buffer were mixed;MTBE was added and the mixture was shaken; the layers were separated via centrifugation; an aliquot was removed from the aqueous layer and reduced involume; the extract was resuspended in methanol and filtered HPLC/MS/ MS LOQ: 62 ng/g ww//
NR Van de Vijver et al., 2005

Fresh water fish: liver Liver was homogenized in water; tetrabutylammonium hydrogen sulfate, (pH10) and Na 2 CO 3 / NaHCO 3 buffer and MTBE were then added; samples wereshaken and centrifuged; an aliquot of the organic layer was transferred to a polypropylenetube; the solution was dried under a gentle nitrogen stream; residue was resuspended in methanol; extract was filtered through nylon mesh filter HPLC/MS/ MS LOD: 72 ng/g ww//
NR Sinclair et al. 2004

Fresh water fish: liver liver homogenized in sodiumcarbonate and TBAS (tetrabutyl-ammonium hydrogen sulfate); the aqueous homogenate was shaken twice for 10 min with MTBE (methyl tert-butyl ether); the combined MTBE extracts were dried in a gentle stream of N 2 ; then water/methanol (1:1) was added and the sample filtered LC/MS/MS LOD: 1.2 ng/g ww//
>80% Moody et al. 2002

212OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 53 Analysis of biota samples Sample Sample preparation Analysis Limit of detectionand/ or quantification// Recovery Reference

Fresh water fishand invertebrates: wholebody homogenates Invertebrate samples required no sample preparation

Fish were homogenized; a small amount of homogenate was weighed and centrifuged in the presence of Na 2 CO 3 , water, ionpairing agent (tetrabutylammonium hydrogensulfate; pH 10), and internal standard, PFHpA (per-fluoroheptanoic acid); for extraction methyl tert-butyl ether was added; organic phasewas separated via centrifugation; the MTBE super-natant was collected; this extraction process was repeated once more and the supernatants combined; the solution was dried under a gentle stream of nitrogen; the analytes were taken up in methanol and the concentrates were filtered LC/MS/MS LOD: 2 ng/g (fish)//
> 68% Martin et al. 2004b

Fish and oyster Ground sample was added todeionised water in a polyethylene tube and mixed; to 1 ml of the mixture TBA solution and Na 2 CO 3 buffer were added; extraction was performed using MTBE; after centrifugation the MTBE layer was separated; the solution was then reduced in volume HPLC/MS LOD: 10 ng/g dw LOQ: 20 ng/g dw// 78-120% Tseng et al., 2006

213OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 53 Analysis of biota samples Sample Sample preparation Analysis Limit of detectionand/ or quantification// Recovery Reference

Mussel and oyster tissue The frozen tissue samples were thawed and homogenized; 1 g of homogenized tissue was trans-ferred to PP centrifugetube and a mixture of perfluorinated com-pounds (external standard) and 30 ml of 0.01 N KOH/methanol were added; the mixture was shaken at ca. 25°C for 16 h; afterwards 1 ml of thesolution was added to a PP bottle containing distilled waterand shaken thoroughly; extraction was carried out using Oasis HLB cartridges; elution of the cartridge was performed via 40% methanol and 100%methanol; after volume reduction the solution was filtered and finally the filter washed with methanol; the eluate was then reduced in a gentle stream of nitrogen HPLC/MS/ MS LOQ: 0.204 ng/g ww//
mean: 100.5% So et al., 2006

Biota (fresh and salt water fish,marine mammals) Sample preparation according to the method of Hansen et al.(2001): samples were homogenized with water; ammoniumcarbonate buffer and tetrabutylammonium hydrogen sulfate (pH 10) were added; extraction of the resulting solution was carried out using tert-butyl methyl ether; extracts were combined and evaporated todryness; the residue was dissolved in methanol and filtered HPLC/MS LOD: 0.4 ng/g ww//
40–72% Berger et al. 2004
Kallenborn et al. 2004

214OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 53 Analysis of biota samples Sample Sample preparation Analysis Limit of detectionand/ or quantification// Recovery Reference

Mammals (bottlenose dolphin) plasma Plasma was placed in a centrifuge tube; internal standard (PFHpA or 13 C-PFOA) Na 2 CO 3and ion-pairing agent (TBAStetrabutylammonium hydrogensulfate pH adjusted to 10) wereadded; the mixture was extracted using MTBE; after centrifugation the organic layer was separated; extracts were blown to dryness with nitrogen; afteraddition of methanol and vortexing, the extract was filtered HPLC/MS/ MS LOQ: 0.5 ng/g//
93% Houde et al., 2005

Marine fish/mammals, and birds: liver and blood Sample preparation acc. to Hansen et al. (2001) HPLC/MS/ MS LOQ: 1–72 ng/g ww (liver) LOQ: 2.5 ng/ml (blood)//<40–90 % (recovery depended onanimals investigated) Kannan et al. 2002b

Marine mammals: blood, liver, and tissue Ion-pair extraction procedure acc. to Hansen et al. (2001); internal standard(1H,1H,2H,2Hperfluorooctanesulfonate and/or perfluorobutane-sulfonate) was added before extraction HPLC/MS/ MS LOQ: 3–20 ng/ml//
75–120% Corsolini and Kannan, 2004

Marine mammals: liver Acc. to Hansen et al. (2001) andHoff et al. (2003): Sample homogenized on ice with water; homogenate was mixed with internal standard (1H,1H,2H,2H-perfluorooctanesulfonic acid, tetrabutylammonium hydrogen sulfate,, sodiumcarbonate buffer; methyl tertbutyl ether was added and the mixture shaken; organic and aqueous layer separation viacentrifugation; after concentrating, the residue was dissolved in methanol and filtered LC/MS/MS LOQ: 10–110 ng/g ww (for all PFCAs)//
NR Van de Vijver et al. 2003 Van de Vijver et al. 2004

215OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 53 Analysis of biota samples Sample Sample preparation Analysis Limit of detectionand/ or quantification// Recovery Reference

Marine food web analysis was conducted according to the method of Hansen etal. (2001) with few modifications HPLC/MS/ MS LOD: 0.2 ng/g// Recovery: NR Tomy et al. 2004

Fish, mussels, amphipods, and algae: tissue Tissue was cut into small pieces; water was added and the samples homogenized; afteraddition of acetonitrile the sample was shaken; the organic and aqueous layers were separated via centrifugation; water was added to the acetonitrile phase; extraction was carried out using SPE cartridges; cartridges were eluted with methanol HPLC/MS/ MS LOQ: 1–10 ng/g ww//
<50% Kannan et al. 2005

Polar bear analysis was conducted according to the method of Hansen etal. (2001): samples were homogenized with sodium carbonate buffer, tetrabutylammonium solution (pH10) and internal standard; extraction with methyl tert-butyl ether; the organic and aqueousphase were separated by centrifugation; MTBE fraction wasevaporated to dryness; the residue was dissolved in methanol/water; the solution was vortexed and filtered LC/MS/MS LOD: 2 ng/g ww//
100% Smithwick et al. 2005b

Polar bear: liver Liver samples were extracted using an ion pairing agent; after drying of the extracts distilled water was added; derivatization with 2,4-difluoroaniline GC/MS NR Ellis et al. 2004

216OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 53 Analysis of biota samples Sample Sample preparation Analysis Limit of detectionand/ or quantification// Recovery Reference

Polar bear: liver Liver tissue homogenized in centrifuge tubes containing Na 2 CO 3 , water, ion-pairingagent TBAS (tetrabutylammonium hydrogensulfate, pH10), and internal standard; extraction with methyl tert-butyl ether; the extraction procedure was repeated several times and the supernatants were collected after centrifugation; solvent was removed by nitrogen stream; addition of water/methanol and vortexing; filtration LC/MS/MS LOD: 2 ng/g ww//
>80% Martin et al. 2004a

Polar bear: liver Homogenized tissue was extracted using an ion-pairingagent and MTBE; extracts combined and evaporated to dryness; residues dissolved in distilled water and acidified to pH 1.0; NaCl, ethyl acetate, 1,3- dicyclohexylcarbodiimide as catalyst and 2,4-difluoroaniline (derivatizing agent) were added; after stirring of the mixture forone hour NaCl was added; mixture separated via a separatory funnel; organic phase was washed with HCl, NaHCO 3 and NaCl solution; organic phase was filtered through Na 2 SO 4 and evaporated to dryness; residueswere dissolved in hexanediethyl ether (95%/5%); further cleanup was performed using a silica gel column and elutionwith hexane-diethyl ether; eluents were evaporated to 200 μL GC-MS LOD: 0.02 pg// Recovery: NR De Silva and Mabury 2004

Beef cattle: blood plasma Plasma was mixed with tetrabutylammonium hydrogen sulfate and buffer (pH 10); the mixture was extracted 3 times withMTBE; after shaking and centrifugation the extract was concentrated under nitrogen; finally, methanol was added and the solution filtered HPLC/MS/ MS NR (solvent blank: 90 pg/ml)//
92% Guruge et al. 2004

217OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 53 Analysis of biota samples Sample Sample preparation Analysis Limit of detectionand/ or quantification// Recovery Reference

Guillemot: eggs Extraction acc. to Hansen et al.(2001): egg samples were homogenized in water; homogenate was mixed with sodiumcarbonate buffer and tetrabutylammonium solution (pH 10); mixture was vortexed and methyl tert-butyl ether added; the organic and aqueous phasewere separated by centrifugation; MTBE fraction was evaporated to dryness; the residue was dissolved in methanol and filtered HPLC/MS/ MS LOD: 3 ng/g//
NR Holmström et al. 2005

Wood mice: liver Analysis was performed according to Giesy and Kannan (2001) HPLC/MS/ MS LOD: 0.11 μg/g ww//
NR Hoff et al., 2004

Turtle: plasma Plasma samples were spikedwith PFBS; plasma, tetrabutylammonium hydrogen sulfate (pH 10) and sodium carbonate were thoroughly mixed and thenshaken with MTBE; after centrifugation, the MBTE extractwas transferred to a polypropylene tube; extract was evaporated under nitrogen; methanolwas added and the mixture vortexed and filtered. HPLC/MS/ MS LOD: 0.05 ng/ml//
70-120% Keller et al. 2004; 2005

Mink and river otters: liver Liver homogenized in water; homogenate was mixed with tetrabutyl-ammonium hydrogen sulfate solution (pH 10) andsodium carbonate buffer; mixture was vortexed and methyltert-butyl ether added; the organic and aqueous phase were separated by centrifugation; MTBE fraction was removedfrom the solution; aqueous solution was rinsed with MTBE and separated twice; rinses were combined; solvent was allowed to evaporate; the residue was dissolved in methanol, vortexed and filtered HPLC/MS/ MS LOQ: 4.5–75 ng/g ww//
53–140% Kannan et al. 2002c

218OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 53 Analysis of biota samples Sample Sample preparation Analysis Limit of detectionand/ or quantification// Recovery Reference

Bald eagle: liver, kidney, muscle, testes, ovaries, and gallbladder Fish: liver, eggs, muscle Frog, mink: liver Snapping turtle: plasma Homogenates were prepared in water; tetrabutylammonium hydrogen sulfate and sodium carbonate buffer (pH 10) were added; mixture was thoroughly shaken; extraction was carried out using methyl tert-butylether; after shaking and centrifugation the MTBE phasewas separated; solvent was removed; the residue was redissolved in methanol and filtered HPLC/MS/ MS LOQ: 7.5–75 ng/g ww (tissue) LOQ: <2.5 ng/ml (plasma)//
<50% Kannan et al. 2005

219OECD SIDS AMMONIUM PERFLUOROO CTANOATE & PERFLUOROOCTANOIC ACID

Table 54 Analysis of other samples

Sample Sample preparation Analysis Limit of detectionand/ or quantification// Recovery Reference

All-weatherjackets for children 10 x 10 cm textile were extracted with ethyl acetate and methanol HPLC/MS LOD: <0.80 μg/m2; LOD: <6.1 ng/g//
NR SNF 2006

Textile (apparel), home furnishingsand carpet samples Samples cut into small pieces;extraction using the corresponding solvent; centrifugation of tubes; LC/MS/MS

. LOQ: 1 ng PFOA/g sample (for samples extracted in waterand sweat simulant)//
100% LOQ: 2.5

– 3 ng/g(for samples extracted in saliva simulant and methanol)// Mawn et al. 2005

PTFE polymer (commercial lot and intermediate of commercial lot) a. Reflux extraction procedure: Five reflux extraction systems run concurrently; solvent was added to each flask; recovery standard was added to one flask; approx. 0.75 g of each polymer was added to three of the flasks; the fifth flask served as solvent blank; flasks were refluxed for two hours, cooledand filtered; solvents were removed
b. Pressurized solvent extraction(PSE) procedure: Five pressurized solvent extractions performed concurrently. Cellsfilled with sand to approximately 3 mm from the top; cells were closed and loaded onto theinstrument; cells were preconditioned; new vessels were set in place; approx. 5/6 of the sand was removed and a weighed quantity of polymer added; cells were refilled to 3 mm from the top with the sand; one vessel for recovery checkand one as solvent blank; extraction was now carried out; solvents were removed LC/MS/MS LOQ: 0.5 ng/g//
89 – 103 % (water,methanol, acetonitrile, ethanol as solvent and reflux conditions)

95 – 99 % (water, methanol, ethanol as solvent and PSE conditions) Larsen et al., 2005

220