ARCHIVED - Environmental Screening Assessment Report on Perfluorooctane Sulfonate

Fate, Exposure and Effects

Environmental Fate of PFOS Precursors

PFOS precursors may be subject to atmospheric transport from their sources to remote areas. While exact transport mechanisms and pathways are currently unknown, the vapour pressures of PFOS precursors, such as N-EtFOSEA and N-MeFOSEA, may exceed 0.5 Pa (1000 times greater than that of PFOS) (Giesy and Kannan 2002). Several PFOS precursors are considered volatile, including N-EtFOSE alcohol, N-MeFOSE alcohol, N-MeFOSA and N-EtFOSA (US EPA OPPT AR226-0620). Two PFOS precursors, N-EtFOSE alcohol and N-MeFOSE alcohol, have been measured in air in Toronto and Long Point, Canada (Martin et al. 2002). For precursors released to the water compartment, the vapour pressure may be significant enough to allow the substance to enter into the atmosphere. For N-EtFOSE alcohol, the tendency to leave the water phase is supported by its relatively high Henry’s law constant (1.9 × 103 Pa•m3/mol) (Hekster et al. 2002). 3M has reported that when these PFOS precursors are present as residuals in products, they could evaporate into the atmosphere when the products containing them are sprayed and dried (US EPA OPPT AR226-0620). The volatility of certain PFOS precursors may lead to their long-range atmospheric transport (Martin et al. 2002). Although evidence of long-range transport of precursors is limited, it is expected that this is at least partially responsible for the ubiquitous presence of PFOS measured at a distance from significant sources.

It is expected that the precursors identified in Appendix 1 will undergo degradation once released to the environment. The perfluorinated moiety is known to be very resistant to degradation, a property attributed to the C-F bond, one of the strongest chemical bonds in nature (~110 kcal/mol) (US EPA OPPT AR226-0547). The perfluorinated chain provides exceptional resistance to thermal and chemical attack (US EPA OPPT AR 226-0547).

Precursors that reach a remote region through the atmosphere or other media are expected to undergo abiotic or biotic degradation to PFOS (Giesy and Kannan 2002; Hekster et al. 2002). The mechanism of this degradation is not well understood; however, it is expected to involve both abiotic and biotic degradation routes. The available experimental environmental degradation rates of PFOS precursors are limited to N-MeFOSE alcohol, N-EtFOSE alchohol, N-MeFOSEA and N-EtFOSEA and are summarized in Table 2.

^a PFOA = perfluorooctanoic acid.
^b N/A = not available.
Source: Hekster et al. (2002).

The two most common intermediate substances used for producing PFOS, N-EtFOSE alcohol and N-MeFOSE alcohol, were tested at several pH concentrations for hydrolysis potential (US EPA OPPT AR226-1030a076, AR226-1030a079). While some of the alcohols disappeared during the test, no PFOS was generated. No hydrolysis studies on N-EtFOSE esters or N-MeFOSE esters were found in a search of US EPA OPPT AR226, which deals entirely with perfluorinated alkyl compounds (US EPA OPPT AR226-0001 through AR226-1040).

The available studies on photolysis show that this transformation mechanism will be of no importance in the breakdown of perfluorinated chemicals. The tests with PFOS, perfluorooctanoic acid (PFOA), POSF and N-EtFOSE alcohol show no photodegradation at all (Hekster et al. 2002; US EPA OPPT AR226-0184, AR226-1030a041). Aqueous photolytic screening studies carried out with N-EtFOSE alcohol, N-MeFOSE alcohol, N-EtFOSA and N-MeFOSA as well as on a surfactant and foamer product showed no direct photolysis, although some underwent indirect photolysis. The primary products were PFOA, perfluorooctane sulfonic acid (PFOSA) and N-EtFOSA (US EPA OPPT AR226-1030a073, AR226-1030a074, AR226-1030a080, AR226-1030a106).

Although experimental evidence on the degradation of PFOS precursors to PFOS is very limited, the precursors are expected to degrade through bacterial-mediated degradation pathways. The biodegradation software, CATABOL, which simulates Organisation for Economic Co-operation and Development (OECD) 302C 28-day biodegradation tests and which has been designed to accommodate perfluorinated compounds, predicts that the majority of those substances identified as precursors (see Appendix 1) will degrade to PFOS (Dimitrov et al. 2004). This degradation has been further supported by expert judgement. It is therefore expected that once those substances listed in Appendix 1 are subjected to a biotic or abiotic degradation mechanism, the perfluorinated moiety that remains will be PFOS. The rate of degradation to PFOS is not considered significant to this assessment, as, over time, these substances are all expected to degrade in the Canadian environment to PFOS.

Environmental Fate of PFOS

Once the precursors degrade to PFOS, the substance will remain indefinitely, as there are no known degradation mechanisms for PFOS in the environment.

Due to the high energy of the C-F bond, PFOS is resistant to hydrolysis, photolysis, aerobic and anaerobic biodegradation and metabolism by vertebrates. The estimated half-life for PFOS is reported as >41 years (Hekster et al. 2002), but may be significantly longer than 41 years. The persistent nature of PFOS is supported by numerous studies (Key et al. 1997; Giesy and Kannan 2002; Hekster et al. 2002; OECD 2002). PFOS is considered to be persistent in the Canadian environment, as the environmental half-life for PFOS is considered to exceed the half-life criteria for persistence as defined by the Persistence and Bioaccumulation Regulations of CEPA 1999 (Government of Canada 2000).

Once PFOS is in the environment, it may enter the food chain or be further distributed at a distance from its source. PFOS has been detected in wildlife at remote sites far from sources or manufacturing facilities (Martin et al. 2004), providing evidence of the bioaccumulation potential and giving support to the persistent nature of PFOS. This also suggests that either PFOS or PFOS precursors may undergo long-range transport. POSF, a precursor and analogue to PFOS, is resistant to atmospheric hydroxyl radical attack and is considered persistent in air, with an atmospheric half-life of 3.7 years (US EPA OPPT AR226-1030a104). In water, PFOS was observed to persist for more than 285 days in microcosms under natural conditions (Boudreau et al. 2003). The OECD PFOS hazard document reviewed several biodegradation studies that indicated that no biodegradation had taken place (OECD 2002).

Examining the physical and chemical properties of PFOS to provide an indication of the environmental fate can be difficult, given the unique physical and chemical characteristics of PFOS. Due to the surface-active properties of PFOS, a log Kow value cannot be determined (OECD 2002). Unlike the situation with other hydrocarbons, hydrophobic and hydrophilic interactions are not the primary partitioning mechanisms, but electrostatic interactions may be more important. It has been suggested that PFOS absorbs via chemisorption (Hekster et al. 2002). A soil adsorption/desorption study using various soil, sediment and sludge matrices found that PFOS appeared to adsorb to all matrices tested (3M Environmental Laboratory 2002). River sediment displayed the most desorption, at 39% after 48 hours, whereas sludge samples did not desorb detectable amounts of test substance. If PFOS does bind to particulate matter in the water column, it can be expected to ultimately settle and reside in sediment.

While the vapour pressure of PFOS is similar to those of other globally distributed compounds (e.g., polychlorinated biphenyls [PCBs], dichlorodiphenyltrichloroethane [DDT]), its water solubility indicates that PFOS is less likely to partition to and be transported in air (Giesy and Kannan 2002). PFOS potassium salt has a water solubility value of 519 mg/L, which has been found to decrease significantly with increasing salt content (US EPA OPPT AR226-0620; Hekster et al. 2002; OECD 2002). Preliminary results of equilibrium partitioning modelling have predicted that PFOS partitions predominantly to the water compartment (80%), with only moderate partitioning to soils and sediments (20%) (US EPA OPPT AR226-0060; CEMC 2001). The OECD review of PFOS data suggested that any PFOS released to a water body would tend to remain in that medium, unless otherwise adsorbed onto particulate matter or taken up by organisms (OECD 2002).

Bioaccumulation

For many organic compounds, the bioaccumulation factor (BAF) may be derived from the octanol/water partition coefficient, because most organic compounds accumulate in lipids. Since perfluorinated surfactants likely elicit a different partitioning behaviour, the Kow is not a suitable predictor for bioaccumulation.

The available, reliable studies on bioaccumulation show that PFOS bioaccumulates and is excreted to a very small extent. Evidence includes calculated BAFs and bioconcentration factors (BCFs), as well as the presence of PFOS in tissue and blood of wildlife in remote areas, including the Canadian Arctic, where manufacturing does not occur.

In an in situ bioaccumulation study following an accidental release of fire-fighting foam into Etobicoke Creek, tissue measurements were taken 6 months after the spill, and water concentrations were taken on day 153 in running water (Moody et al. 2002). Calculated BAFs in fish ranged from 6300 to 125 000 for PFOS, based on measured concentrations in common shiner (Notropus cornutus) liver and surface water. This BAF is high in comparison with BCF values available. Moody et al. (2002) suggested that accumulated perfluorinated derivatives are metabolized to PFOS, thus overestimating the BAF of PFOS. Nevertheless, it is evidence of the bioaccumulative nature of PFOS, and available BAFs remain substantially higher than the bioaccumulation criterion of 5000 under the Persistence and Bioaccumulation Regulations of CEPA 1999 (Government of Canada 2000).

Giesy and Kannan (2002) determined the water concentration and whole fish burdens of PFOS at Guntersville Dam in Alabama from data supplied in US EPA OPPT AR226-1030a161. From these values, BAFs ranging from 830 to 26 000 were calculated for PFOS for channel catfish (Ictalurus punctatus) and largemouth bass (Micropterus salmoides), respectively (Purdy 2002b). The difference could be due to largemouth bass being higher in the food chain and receiving more PFOS via food (US EPA OPPT AR226-1030a161).

PFOS has been found to bioconcentrate in fish (OECD 2002). Estimated BCFs of 1100 (carcass), 5400 (liver) and 4300 (blood) have been reported for juvenile rainbow trout (Oncorhynchus mykiss) (Martin et al. 2003a). In fish livers collected from 23 different species in Japan, BCFs were calculated to range from 274 to 41 600 (mean 5500) (Taniyasu et al. 2003). While fish may be able to eliminate PFOS via their gills, this mode of elimination is not available to higher trophic level predators (Martin et al. 2003b), and high concentrations of PFOS have been found in the liver and blood of higher trophic level predators that consume fish (e.g., polar bears, mink and birds). Indeed, maximum levels of PFOS in liver of Canadian Arctic biota have been reported for mink (20 µg/kg), trout (50 µg/kg), seal (37 µg/kg), fox (1400 µg/kg) and polar bear (>4000 µg/kg) (Martin et al. 2003b) (see Table 3). Species differences for the elimination half-life of PFOS in biota have been shown to vary significantly: 15 days (fish), 100 days (rats) and 200 days (monkeys) (OECD 2002; Martin et al. 2003a). In addition to information on PFOS, the US Interagency Testing Committee estimated BCFs for N-EtFOSEA and N-MeFOSEA using structure-activity models to be 5543 and 26 000, respectively (Giesy and Kannan 2002).

Exposed rodents have shown preferential distribution of perfluorinated alkyl compounds in blood and liver rather than in lipids (Taniyasu et al. 2002; Martin et al. 2003b). BCFs for perfluorinated alkyl compounds have also been found to be higher in fish blood and liver than in fish carcass (Martin et al. 2003a).

Some PFOS precursors have been measured in air (Martin et al. 2002). These precursors, N-MeFOSE alcohol and N-EtFOSE alcohol, are relatively volatile, especially for such large chemicals, and they have relatively high octanol/water partition coefficients. They could be entering food chains by partitioning into biota and then undergoing degradation to PFOS somewhere along the food chain. The amount of PFOS and precursors in an animal depends on what it is eating, how much the prey metabolizes intermediates and which degradation pathways occur in the predator (Purdy 2002b).

When rats metabolize N-MeFOSE-based compounds, several metabolites have been confirmed in tissue samples, including PFOS and N-MeFOSE alcohol (3M Environmental Laboratory 2001a, 2001b). PFOS appears to be the final product of rat and probably other vertebrate metabolism of POSF-based substances.

Environmental Concentrations

Martin et al. (2002) measured the air in Toronto and Long Point for some precursors of PFOS. They found an average N-MeFOSE alcohol concentration of 101 pg/m3 in Toronto air and 35 pg/m3 at Long Point. The average concentrations of N-EtFOSE alcohol were 205 and 76 pg/m3, respectively. No air concentration data for PFOS or precursors from other countries were found.

In June 2000, PFOS was detected in surface water as a result of a spill of a fire-fighting foam from the Toronto International airport into nearby Etobicoke Creek. Concentrations of PFOS ranging from <0.017 to 2210 µg/L were detected in creek water samples over a 153-day sampling period. PFOS was not detected at the upstream sample site (Moody et al. 2002). No Canadian monitoring data for PFOS were found in sediment, effluent or sludge.

US data for PFOS are available from one study of six cities. PFOS was detected in quiet water (i.e., a pond) (2.93 µg/L) and sewage treatment effluent (0.048-0.45 µg/L) and sludge (60.2-130 µg/kg dry sludge) at cities (Port St. Lucie, Florida, and Cleveland, Tennessee) with no significant fluorochemical activities (US EPA OPPT AR226-1030a111). PFOS was also detected in drinking water (0.042-0.062 µg/L), surface water (not detected [n.d.] to 0.08 µg/L), sediments (n.d.-0.78 µg/kg dry sediment), sewage treatment effluents (0.04-5.29 µg/L) and sludge (57.7-3120 µg/kg) and landfill leachate (n.d.-53.1 µg/L) of four cities that have manufacturing or industrial use of fluorochemicals. Detection limits were 0.0025 µg/L for water and 0.080 µg/kg wet weight for sediment and sludge. Sediment concentrations appear to be approximately 10-fold higher than water concentrations, indicating that there is a tendency to partition from the water to sediment.

Samples of the surface microlayer of natural water were also collected but not analyzed in the six-city US study. The reason given was that an accepted method for microlayer sample handling was not available, and consequently the evaluation and interpretation of the data were not available (US EPA OPPT AR226-1030a111). It is predicted that these samples would have shown high concentrations, because PFOS is surface active and has been found, in a limited sample set, to magnify 200-fold in the surface microlayer compared with the concentration in the underlying water (Purdy 2002b).

In a recent monitoring study near the vicinity of a fluorochemical manufacturing facility located on the Tennessee River (Alabama), PFOS was detected in all surface water and sediment samples collected. The highest concentrations for surface water (151 µg/L) and sediment (5930 µg/kg wet weight; 12 600 µg/kg dry weight) were found at a location near the point of discharge of a combined industrial effluent. However, the study found that downstream concentrations were not statistically greater than those upstream and concluded that the combined industrial effluent did not significantly affect fluorochemical (including PFOS) concentrations in the main stem of the river. For the upstream reference site (Guntersville Dam), estimated average PFOS surface water and sediment concentrations were 0.009 µg/L and 0.18 µg/kg, respectively (US EPA OPPT AR226-1030a161).

In another study, low levels of PFOS were found throughout a 130-km stretch of the Tennessee River (Hansen et al. 2002). The average PFOS concentration upstream of the fluorochemical manufacturing facility was 0.032 µg/L, suggesting an unidentified source of PFOS entering the river upstream.

Table 3 presents the levels of PFOS found in wildlife worldwide. A recent Canadian survey detected PFOS and other perfluorinated acids in fish, birds and mammals from various locations in the Canadian Arctic (Martin et al. 2004). Data are also available for a variety of species, including oysters along the Gulf of Mexico and southern Atlantic coast of the United States, fish-eating birds in Asia, Europe and North America, seals in the Caspian Sea, and polar bears and mink in North America.

^a References: US EPA OPPT AR226-1030a156, AR226-1030a157, AR226-1030a158, AR226-1030a159, AR226-1030a160 as summarized by Giesy and Kannan (2002); except entries denoted by *, which are from Martin et al. (2004), Kannan et al. (2002) and Giesy (2003).
^b Units are parts per billion (ppb) = µg/kg for tissue; µg/L for fluids.
^c n.r. = not reported.

The highest tissue concentration in Table 3 is 4870 µg/kg in mink liver from the Midwestern United States. In Canada, the highest PFOS concentration was found in polar bear liver (maximum = >4000 µg/kg, mean = 3100 µg/kg; n = 7) (Martin et al. 2004). The PFOS concentrations in polar bear liver were higher than any other previously reported concentrations of persistent organochlorine chemicals (e.g., PCBs, chlordane, hexachlorocyclohexane) in polar bear fat. A general data trend indicated that mammals feeding at higher trophic levels had higher PFOS concentrations than those feeding at lower trophic levels. Elsewhere, PFOS concentrations in plaice (Pleuronectes platessa) liver (7760 µg/kg) from the Western Scheldt estuary (southwestern Netherlands) and ornate jobfish (Pristipomoides argyrogrammicus) liver (7900 µg/kg) from Kin Bay (Japan) are among the highest PFOS concentrations ever reported in wildlife (fish) (Hoff et al. 2003; Taniyasu et al. 2003). Factors helping to explain such high concentrations may be the proximity of a PFOS manufacturing plant (upstream of estuary) and an army base (Kin Bay, Japan) that could be using PFOS in fire-fighting operations.

Effects

The toxicity of PFOS has been studied in a variety of aquatic and terrestrial species, including aquatic plants, invertebrates and vertebrates and terrestrial invertebrates, birds and mammals. Adverse effects range from growth inhibition, histopathological effects, atrophied thymus, change in species diversity in a microcosm and mortality. Toxicity data are essentially limited to PFOS. The following is a summary of the key studies used to identify the Critical Toxicity Value (CTV) for PFOS. A more complete review of effects is given in the OECD hazard review of PFOS, which discusses effects on fish, invertebrates, aquatic plants (algae and higher plants), amphibians and microorganisms (OECD 2002). Additional studies by Boudreau et al. (2003) and Sanderson et al. (2002) not available in OECD (2002) are also summarized.

The most sensitive endpoint in aquatic organisms occurred in a flow-through bioconcentration study with bluegill (Lepomis macrochirus) using PFOS potassium salt. No significant mortality was seen at an exposure concentration of 0.086 mg/L over a 62-day uptake phase; however, significant mortality was observed after a 35-day exposure to 0.87 mg/L. The study was stopped because all the fish either had died or had been sampled (US EPA OPPT AR226-1030a042). The No-Observed-Effect Concentration (NOEC) of 0.086 mg/L is the lowest no-adverse-effect concentration for aquatic organisms and was therefore selected as the CTV for aquatic organisms.

Results have been published from a laboratory evaluation of the toxicity of PFOS to five aquatic organisms: green algae (S. capricornutum and C. vulgaris), duckweed (L. gibba) and water flea (D. magna and D. pulicaria) (Boudreau et al. 2002). NOEC values were generated from the most sensitive endpoints for all organisms. Based on effect (immobility) values, the most sensitive of the organisms in this study was D. magna, with a 48-hour immobility NOEC of 0.8 mg/L; the accompanying LC₅₀ was 112 mg/L, and the 48-hour IC₅₀ for growth inhibition was 130 mg/L. The 21-day NOEC for lethality for D. magna was 5.3 mg/L. Autotroph inhibition of growth NOEC values were 5.3 mg/L, 6.6 mg/L and 8.2 mg/L for S. capricornutum, L. gibba and C. vulgaris, respectively.

In an aquatic microcosm study (Boudreau et al. 2003), a field evaluation assessed the toxicological risk associated with PFOS across levels of biological organization. The zooplankton community was significantly affected by the treatment for all sampling times. A community-level NOEC of 3.0 mg/L was determined for the 35-day study. The most sensitive taxonomic groups, Cladocera and Copepoda, were virtually eliminated in the 30 mg/L treatments after 7 days, although specific survival rates were not quantified.

In a laboratory microcosm study that examined impacts to zooplankton following exposure to PFOS, adverse effects were observed at 10 mg/L over 14 days; several species were significantly reduced or eliminated (Sanderson et al. 2002). In comparison with controls, exposures of 10 mg/L and 30 mg/L resulted in an average 70% change in species diversity and total zooplankton. The most sensitive species in the study was Cyclops diaptomus. The statistically significant effect concentrations for all species endpoints (abundance) were above 1 mg/L.

A fathead minnow (Pimephales promelas) embryo-juvenile flow-through chronic study determined a NOEC of 0.3 mg/L over a 42-day exposure period. This value was for both survival and growth (US EPA OPPT AR226-0097). The slightly higher NOEC may be due to the shorter exposure time. In acute tests, the lowest 96-hour LC₅₀ for freshwater fish species was 4.7 mg/L for the fathead minnow (P. promelas); in salt water, a 96-hour LC₅₀ of 13.7 mg/L was reported for rainbow trout (O. mykiss) (OECD 2002). In a 96-hour static acute study using the freshwater mussel, Unio complamatus, the NOEC for mortality was 20 mg/L and the LC₅₀ was 59 mg/L (US EPA OPPT AR226-0091, AR226-1030a047). The most sensitive saltwater invertebrate studied was the saltwater mysid, Mysidopsis bahia. Survival, growth and reproduction were assessed over an exposure period of 35 days. The NOECs determined for growth and reproduction were both 0.25 mg/L (US EPA OPPT AR226-0101). In acute toxicity testing, a 96-hour LC₅₀ of 3.6 mg/L was reported for the mysid shrimp (OECD 2002). There was one study reported for embryo teratogenesis in aquatic organisms, which involved a 96-hour static renewal study on the frog, Xenopus laevis (US EPA OPPT AR226-1030a057). The minimum concentration that inhibited growth was 7.97 mg/L. The LC₅₀ for mortality was 13.8 mg/L, the EC₅₀ for malformed embryos was 12.1 mg/L and the NOEC for embryo malformation was 5.2 mg/L. Calculated teratogenic indices ranged from 0.9 to 1.1, indicating that PFOS has a low potential to be a developmental hazard in this species.

PFOS is toxic to birds. Acute dietary studies were conducted on mallard (Anas platyrhynchos) and northern bobwhite (Colinus virginianus) (US EPA OPPT AR226-1030a049). Birds were fed PFOS in the diet for 5 days. Mortality, body weight and food consumption were monitored throughout. PFOS levels were quantified in sera and liver of mallards and bobwhites sacrificed on days 8 and 22 post-exposure, as well as from some animals that died before the scheduled sampling times. Mallard was the more sensitive of the two species tested, and the most sensitive endpoint was the 8-day Lowest-Observed-Effect Concentration (LOEC) for reduced body weight gain, at 29.7 mg PFOS/kg liver wet weight. This value is the CTV for birds (NOEC = 15.3 mg PFOS/kg liver wet weight). For northern bobwhite, the mean LOEC in the day 8 group was 70.3 mg/kg liver wet weight, and the NOEC was 45.2 mg/kg liver wet weight.

As no wild mammal studies were found, laboratory mammal studies were used as surrogates for wild mammals. The CTV for mammal (liver) and bird (serum) was selected from a 2-year dietary rat study in which histopathological effects in the liver were seen in males and females at intakes as low as 0.06-0.23 mg PFOS/kg bw per day and 0.07-0.21 mg PFOS/kg bw per day, respectively (Covance Laboratories, Inc. 2002). Average values were determined for males and females, to establish Lowest-Observed-Effect Levels (LOELs) of 40.8 mg/kg in liver and 13.9 mg/L in serum.

Supporting evidence for a NOEC in the low mg/kg or mg/L range in liver and sera includes results from a two-generation rat study (US EPA OPPT AR226-0569). In this study, the NOECs were determined to be 0.1 mg/kg bw per day dosed via gavage, 5.3 mg/L in sera and 14.4 mg/kg in liver. The LOECs were 0.4 mg/kg bw per day, 19 mg/L sera and 58 mg/kg liver. The effect was reduction in dam body mass (US EPA OPPT AR226-0569).

Additional studies in primates are summarized in the screening assessment report for PFOS and its precursors prepared by Health Canada (2004).

The OECD review summarizes data indicating moderate to high toxicity of PFOS to honey bees (Apis mellifera). In an acute oral test, a 72-hour LD50 for ingestion of PFOS was 0.40 µg/bee, and a 72-hour No-Observed-Effect Level (NOEL) was 0.21 µg/bee. A contact test found a 96-hour LD50 of 4.78 µg/bee and a 96-hour NOEL of 1.93 µg/bee.

Results have been reported for an acute toxicity study with the earthworm in an artificial soil substrate (US EPA OPPT AR226-1106). The PFOS potassium salt 14-day LC₅₀ was determined to be 373 mg/kg bw, with a 95% confidence interval of 316-440 mg/kg bw. The 14-day NOEC for burrowing behaviour, body weight and clinical signs of toxicity was 77 mg/kg bw, and the 14-day LOEC for the same endpoints was 141 mg/kg bw.

Body mass reduction or poor food efficiency was seen in most toxicity studies and species (Haughom and Spydevold 1992; Campbell et al. 1993a, 1993b; US EPA OPPT AR226-0137, AR226-0139, AR226-0144, AR226-0949, AR226-0953, AR226-0956, AR226-0957, AR226-0958, AR226-0967). This is consistent with the mechanism of toxicity being the uncoupling of oxidative phosphorylation (US EPA OPPT AR226-0167, AR226-0169, AR226-0240). This mode of action, however, is not known with certainty to explain PFOS toxicity. There are other mechanisms that can be hypothesized. A study with rats (Luebker et al. 2002) tested the hypothesis that PFOS, PFOA and other perfluorinated chemicals can interfere with the binding affinity and capacity of liver binding proteins for fatty acids; the results revealed that the most potent competitor is PFOS. A study with common carp (Cyprinus carpio) by Hoff et al. (2003) has suggested that PFOS induces inflammation-independent enzyme leakage through liver cell membranes that might be related to cell necrosis. It was also suggested that PFOS might interfere with homeostasis of DNA metabolism.