Draft Screening Assessment Perfluorooctanoic Acid, its Salts, and its Precursors

Environment Canada
Health Canada

October 2010

Under the Canadian Environmental Protection Act, 1999 (CEPA 1999), the Ministers of the Environment and of Health have conducted a screening assessment of perfluorooctanoic acid (PFOA), Chemical Abstracts Service Registry Number 335-67-1, its salts and its precursors under sections 68 and 74 of CEPA 1999. Some of these substances were categorized in under section 73 of CEPA 1999. PFOA was also assessed due to its persistent nature, widespread occurrence in biota, presence in the Canadian Arctic due to long range transport, and international interest in emerging science indicating a potential concern for the environment and human health from PFOA and its salts. In addition, precursors to PFOA were considered in this assessment on the basis of their contribution to the total presence of PFOA and its salts in the environment.

PFOA is an anthropogenic substance belonging to a class of chemicals known as perfluorocarboxylic acids (PFCAs). PFCAs, in turn, belong to the broader class of chemicals known as perfluoroalkyls (PFAs). The term “PFOA” may refer to the acid, its conjugate base or its principal salt forms. Historical uses of PFOA include applications in industrial processes and in commercial and consumer products. It continues to be used as a reactive intermediate, with its salts used as processing aids in the production of fluoropolymers and fluoroelastomers. PFOA is not manufactured in Canada; however, quantities of the ammonium salt are imported.

Environment

PFOA may be found in the environment due to releases from fluoropolymer manufacturing or processing facilities, effluent releases from wastewater treatment plants, landfill leachates and due to degradation/transformation of PFOA precursors. Such precursors may include parent compounds, chemical products containing PFOA (either as part of formulations or as unintended residuals) and substances transforming to intermediates that ultimately degrade to PFOA. Potential precursors also include related fluorochemicals (e.g., fluorotelomer alcohols [FTOHs], fluorotelomer iodides and fluorotelomer olefins), some of which are currently used and detectable in the atmosphere and can degrade or transform to PFOA through biotic or abiotic pathways.

Once in the environment, PFOA is persistent and not known to undergo any further abiotic or biotic degradation under relevant environmental conditions. PFOA is highly water soluble and typically present as an anion (conjugate base) in solution. It has low vapour pressure; therefore, the aquatic environment is expected to be its primary sink, with some additional partitioning to sediment. The presence of PFOA in the Canadian Arctic indicates the long-range transport of PFOA (e.g., via ocean currents) or of volatile precursors to PFOA (e.g., via atmospheric transport).

PFOA has been detected at trace levels in the northern hemisphere. In North America, higher levels were measured in surface waters in the vicinity of US fluoropolymer manufacturing facilities (<0.025–1900 µg/L) and in groundwater near US military bases (not detected [ND] to 6570 µg/L). PFOA was detected in effluent from Canadian wastewater treatment facilities at concentrations ranging from 0.007 to 0.055 µg/L. PFOA was also detected in the influent at US wastewater treatment facilities at concentrations ranging from 0.0074––0.089 µg/L.

Trace levels of PFOA have been measured in Canadian fresh water (ND–11.3 µg/L) and freshwater sediments (0.3–7.5µg/kg). PFOA has also been detected in a variety of Canadian biota (ND–90 µg/kg wet weight [kg-ww] tissue) in southern Ontario and the Canadian Arctic. The highest concentration of PFOA in Canadian organisms was found in the benthic invertebrate Diporeia hoyi at 90 µg/kg-ww, followed by burbot liver at 26.5 µg/kg-ww, polar bear liver at 13 µg/kg-ww, caribou liver at 12.2 µg/kg-ww, ringed seal liver at 8.7 µg/kg-ww and walrus liver at 5.8 µg/kg-ww. Following a spill of fire-fighting foam in Etobicoke Creek (Ontario), PFOA was measured in common shiner liver at a maximum of 91 µg/kg-ww. However, current PFOA concentrations in Canadian biota (tissue specific and whole body) are below the highest concentration found in US biota (up to 1934.5 µg/kg-wwin gar liver).

Temporal or spatial trends in PFOA concentrations in guillemot eggs, lake trout, thick-billed murres, northern fulmars or ringed seals could not be determined. However, temporal trends were found for PFOA concentrations in polar bears and sea otters. PFOA doubling time in liver tissue was calculated to be 7.3 ± 2.8 years for Baffin Island polar bears and 13.9 ± 14.2 years for Barrow, Alaska, polar bears; central East Greenland polar bears showed an annual increase of 2.3% in PFOA concentrations. Concentrations of PFOA also increased significantly over a 10-year period for adult female sea otters.

Unlike other organic pollutants that are persistent and found in biota, certain perfluorinated substances, such as PFOA, are present mainly as ions in environmental media. Due to the perfluorination, the hydrocarbon chains are oleophilic and hydrophobic and the perfluorinated chains are both oleophobic and hydrophobic. PFOA primarily binds to proteins in biota and preferentially partitions to liver, blood and kidney rather than to lipid tissue. The numeric criteria for bioaccumulation, outlined in the Persistence and Bioaccumulation Regulations of CEPA 1999, are based on bioaccumulation data for aquatic species (fish) only and for substances that preferentially partition to lipids. As a result, the criteria do not account for the bioaccumulation of PFOA that is preferentially partitioning in the proteins of liver, blood and kidney in terrestrial and marine mammals.There is experimental evidence indicating that PFOA is not highly bioaccumulative in fish. Reported laboratory bioconcentration factors for fish species (primarily rainbow trout) ranged from 3.1–27. In the pelagic aquatic food web of Lake Ontario, two studies indicate that PFOA concentrations do not biomagnify with increasing trophic level. However, these results should not be extrapolated to other species, since gills provide an additional mode of elimination for PFOA that air-breathing organisms, such as terrestrial and marine mammals, do not possess. Field studies indicating biomagnification factors greater than 1 for Arctic and other mammals (such as narwhal, beluga, polar bear, walrus, bottlenose dolphins, and harbour seals) suggest that PFOA may bioaccumulate and biomagnify in terrestrial and marine mammals. Reported field biomagnification factors for terrestrial and marine mammals ranged from 0.03–31. Polar bears, as the apex predator in the Arctic marine food web, have been shown to be the most contaminated with PFOA relative to other Arctic terrestrial organisms.

In traditional toxicity studies, PFOA exhibits moderate to low acute toxicities in pelagic organisms, including fish (70–2470 mg/L). PFOA exhibits low chronic toxicities in benthic organisms (>100 mg/L). There is one study on the toxicity of PFOA and its salts in avian wildlife. In this study, PFOA was found to have no effect on embryonic pipping success for white leghorn chickens at concentrations up to 10 µg/g. However, PFOA accumulated in the liver of these embryos to concentrations greater than the initial whole-egg concentration

There are studies showing the potential for PFOA to affect endocrine function where visible effects may not be apparent until the organisms reach adulthood. In female and male rare minnows, 3–30 mg/L PFOA elicited inhibition of the thyroid hormone biosynthesis genes, induced vitellogenin expression in males, developed oocytes in the testes of male fish and caused ovary degeneration in females.

There are other studies showing hepatotoxicity, immunotoxicity, and chemosensitivity. For example, a PFOA concentration of 0.02 µg/L increased the chemosensitivity in marine mussels. PFOA at 25.9 mg/L, activated the mammalian peroxisome proliferator–activated receptor a (PPARa) in the livers of Baikal seals--PPARa plays a critical physiological role as a lipid sensor and a regulator of lipid metabolism. Field data also reveal that there may be increases in indicators of inflammation and immunity in bottlenose dolphins related to PFOA concentrations, suggesting possible autoimmune effects. Another field study has also suggested that low levels of PFOA may alter biomarkers of health in loggerhead sea turtles.

PFOA is persistent in all media in the environment and can bioaccumulate and biomagnify in terrestrial and marine mammals. Given these inherent properties of PFOA, together with environmental concentrations that may approach effect levels affecting endocrine function (including vitellogenin induction, feminization in male fish, ovary degeneration in female fish and hepatotoxicity), current temporal trends of PFOA in polar bears, the widespread occurrence of PFOA in biota, including in remote areas, and the fact that other PFCAs and precursors to PFOA may contribute to the overall additive or synergistic impact of PFOA in biota, it is proposed that PFOA, its salts and its precursors are entering the environment in a quantity or concentration or under conditions that have or may have an immediate or long-term harmful effect on the environment or its biological diversity.

Human Health

In humans, PFOA is well absorbed by all routes of exposure; it has not been demonstrated to be metabolized and has a relatively long half-life. Salts of PFOA are expected to dissociate in biological media to produce the perfluorooctanoate (PFO) moiety, and are therefore considered toxicologically equivalent to PFOA. Low concentrations of PFOA have been identified in blood samples from non-occupationally exposed Canadians, including newborns, indicating environmental exposure to PFOA and/or compounds that can degrade to PFOA. The available data indicate that Canadians are exposed to PFOA and its precursors in the environment, including via air, drinking water and food; and from the use of consumer products, such as new non-stick cookware and perfluorinated compound (PFC)-treated apparel and household materials such as carpets and upholstery. Canadians are also potentially exposed to PFOA in utero and through lactational transfer. The relative contributions of PFOA and its salts and precursors to total PFOA exposure were not characterized; rather the focus was on aggregate exposure to the moiety of toxicological concern, PFOA.

Epidemiological studies have not identified a causal relationship between PFOA exposure and adverse health effects in humans. Therefore, toxicity studies in laboratory animals were used to determine the critical effects and associated serum levels of PFOA. Following oral dosing of PFOA ammonium salt (APFO), increased liver weight in mice and altered lipid parameters in rats were observed in short-term (14-day) toxicity studies; increased liver weight was noted in a 26-week toxicity study in monkeys; and increased liver weight in dams, alterations in fetal ossification and early puberty in male pups were found in a developmental toxicity study in mice.

In 2-year carcinogenicity bioassays in rats, males administered a high dose of APFO in the diet had significantly higher incidences of adenomas of the liver hepatocytes, Leydig cells in the testes and pancreatic acinar cells. No evidence of carcinogenic activity was seen in the female rats. Liver tumours in male rats may be induced via liver toxicity resulting from PFOA-induced peroxisome proliferation, and additional pathways secondary to peroxisome proliferation may be involved in the generation of tumours at other sites. As primates are much less suspectible than rodents to peroxisome proliferation, the PFOA-induced tumours in male rats are considered to have little or no relevance for humans. Although blood levels of PFOA were not determined in the chronic studies, the oral dose of APFO was several times higher than those in the critical short-term and subchronic studies. Although there is some evidence to suggest that PFOA may be capable of causing indirect oxidative DNA damage, the genotoxicity database indicates that PFOA is not mutagenic. Thus, as the tumours observed in male rats are not considered to have resulted from direct interaction with genetic material, a threshold approach is used to assess risk to human health.

The assessment of PFOA is based on a comparison of the margin between the levels of PFOA in the blood (serum) of humans and serum levels that are associated with the development of adverse effects in laboratory animals. This approach aggregates exposure to PFOA from all sources, including those resulting from releases from fluoropolymer manufacturing or processing facilities, effluent releases from sewage treatment plants, landfill effluents, or degradation/transformation of PFOA precursors.

Comparison of the PFOA serum levels associated with adverse effects in laboratory animals (13–77 µg/mL) with the serum levels found in non-occupationally exposed adults and in children (0.0034–0.010 µg/mL) results in margins of exposure of =1300. These margins are considered to be adequately protective to account for uncertainties in the hazard and exposure databases.

Based on the available information on the potential to cause harm to human health and the resulting margins of exposure, it is proposed that PFOA and its salts are not entering the environment in a quantity or concentration or under conditions that constitute or may constitute a danger in Canada to human life or health. Precursors of PFOA were not individually assessed, but were considered in terms of their contribution to total PFOA exposure because they can degrade to PFOA in the environment.

Conclusion

Based on available information for environmental and human health considerations, it is proposed to conclude that PFOA, its salts and its precursors meet one or more of the criteria set out in section 64 of CEPA 1999.

PFOA is highly persistent in the environment and meets the criteria for persistence as defined in the Persistence and Bioaccumulation Regulations. While there is scientific evidence that PFOA and its salts can accumulate and biomagnify in terrestrial and marine mammals, PFOA and its salts do not meet the criteria for bioaccumulation as defined in the Persistence and Bioaccumulation Regulations.

Where relevant, research and monitoring will support verification of assumptions used during the screening assessment and, where appropriate, the performance of potential control measures identified during the risk management phase.

This screening assessment was conducted pursuant to sections 68 and 74 of the Canadian Environmental Protection Act, 1999 (CEPA 1999) (Canada 1999).

A screening assessment was undertaken on perfluorooctanoic acid (PFOA) (Chemical Abstracts Service Registry Number 335-67-1) and its salts. In addition, precursors to PFOA were considered on the basis of their contribution to the total presence of PFOA and its salts. The ammonium salt, CAS RN 3825-26-1, and the precursors, CAS RN 53515-73-4, CAS RN 678-39-7, CAS RN 65530-61-2, and CAS RN 70969-47-0 were found to meet the ecological categorization criteria for persistence, bioaccumulation potential and inherent toxicity to non-human organisms. However, none of these substances were considered to be a high priority for assessment of potential risks to human health, based upon application of the simple exposure and hazard tools developed by Health Canada for categorization of substances on the Domestic Substances List.

Screening assessments focus on information critical to determining whether a substance meets the criteria as set out in section 64 of CEPA 1999. Screening assessments examine scientific information and develop conclusions by applying a weight of evidence approach and precaution

This screening assessment includes consideration of information on chemical properties, hazards, uses and exposure. Data relevant to the screening assessment of these substances were identified in original literature, review and assessment documents and stakeholder research reports and from recent literature searches, up to November 2009 for ecological sections of the document and February 2009 for human health sections of the document. During the peer review process, an additional literature search was carried out to update the coverage of the human health data to November 2009. In addition, an industry survey on perfluoroalkyls/fluoroalkyls was conducted for the years 2000 and 2004 through Canada Gazette Notices issued pursuant to section 71 of CEPA 1999 (Canada 1999; Canada 2000b, 2004). These surveys collected data on the Canadian manufacture, import, uses and releases of perfluoroalkyls/fluoroalkyls. Toxicological studies were also submitted by industry under section 70 of CEPA 1999.

The approach taken in the ecological screening assessment is to examine various supporting information and develop conclusions based on multiple lines of evidence, such as persistence, exposure, trends, inherent toxicity, bioaccumulation and widespread occurrence in the environment. Evaluation of risk to human health involves consideration of data relevant to estimation of exposure (non-occupational) of the general population as well as information on health hazards. Decisions for human health are based on the nature of the critical effect and/or margins between conservative effect levels and estimates of exposure, taking into account confidence in the completeness of the identified databases on both exposure and effects, within a screening context. The screening assessment does not present an exhaustive or critical review of all available data. Instead, it presents the critical studies and lines of evidence supporting the conclusions^[1].

This draft screening assessment was prepared by staff in the Existing Substances programs at Health Canada and Environment Canada. The draft ecological assessment has undergone external written peer review/consultation. The draft human health screening assessment was externally reviewed by staff of Toxicology Advice & Consulting Limited; Dr. Sean Hayes (Summit Toxicology); Dr. Greg Kedderis (private consultant); Dr. Kannan Krishan (University of Montreal); and Dr. Donna Vorhees (Science Collaborative) for adequacy of data coverage and defensibility of the conclusions. Statements in this document do not necessarily reflect the opinions of the reviewers. However, all reviewer comments were considered carefully and, where appropriate, addressed. The final content and outcome of the draft screening risk assessment remain the responsibility of Health Canada and Environment Canada.

The critical information and considerations upon which the draft assessment is based are summarized below.

Footnote

Information on the identity of perfluorooctanoic acid (PFOA) is presented in Table 1. PFOA is an anthropogenic compound with a chain length of eight carbons, seven of which are perfluorinated. It belongs to the broad class of chemicals known as perfluorocarboxylic acids (PFCAs), which, in turn, belong to the broader class of chemicals known as perfluoroalkyls (PFAs). The term “PFOA” may refer to the acid, its conjugate base or its principal salt forms (Table 2). The term PFOA is not interchangeable with commercial mixtures containing PFOA, as these mixtures are often not well characterized and could include any product that contains even a small amount of PFOA. PFOA may also be referred to as C8, as well as by other synonyms or trade names. A more detailed discussion of the identity, nomenclature and trade names of PFOA is provided in Ellis et al. (2004b). The most common commercially used salt form of PFOA is the ammonium salt, referred to as APFO (see chemical structure in Table 1).

Table 1. Substance Identity

Environment Canada considered perfluoroalkyl compounds on the basis of expert judgement, chemical structures and biodegradation estimation modelling using CATABOL (c2004–2008) (Mekenyan et al. 2002). Using these approaches, the structures were analyzed for their potential to degrade to PFOA. CATABOL (c2004–2008) was trained on the basis of the Japanese Ministry of International Trade and Industry (MITI) biodegradation test results and predicts biodegradation over a period of 28 days. It is acknowledged that due to the limited perfluorinated degradation data in the training set, some degradation products generated by CATABOL (c2004–2008) may be of limited reliability. It should be noted that the degradation process will be longer for perfluorinated chemicals, but it is difficult to estimate how much longer, especially for high-molecular-weight substances such as oligomers and polymers.

Precursors to PFOA (Table 2) have been considered in this assessment; however, this list should not be considered exhaustive. A more complete list can be found under the Notice with respect to certain perfluoroalkyl and fluoroalkyl substances at: http://www.gazette.gc.ca/archives/p1/2005/2005-01-15/html/notice-avis-eng.html. It should also be noted that some precursors to PFOA may also be considered as precursors to long-chain (C9-C20) perfluorocarboxylic acids. This assessment defines precursors as substances where the perfluorinated alkyl moiety has the formula C_nF_2n+1 (where n= 7 or 8) and is directly bonded to any chemical moiety other than a fluorine, chlorine or bromine atom.

Table 2. List of PFOA and its Principal Salts and Precursors¹

Water solubility is a key property affecting PFOA and its salts in the environment. The free acid and the ammonium salt are solid at 20°C. The free acid readily dissociates to perfluorooctanoate (PFO), the conjugate base that is commonly measured in environmental media and biological samples. The conjugate base of the PFOA acid and its salts is highly soluble in water. It has been noted that PFOA salts self-associate at the surface but disperse with agitation and form micelles at higher concentrations (US EPA 2003).

Solubility is dependent on the acid dissociation constant (pK_a) of the acid form, and the pK_a value commonly reported and used is approximately 2.5 (Kissa 1994). However, Burns et al. (2008) determined a pK_a of 3.8 ± 0.1, suggesting that the neutral species can exist in the environment. Jasinski et al. (2009) determined a pK_a of 3.3 ± 0.4. Goss (2008), using analogues and molecular models, suggested that the pK_a is more likely to be close to -0.5. Goss (2008) suggested that PFOA is expected to have a low pK_a, such that 99% of the compound will occur in its anionic form (i.e., PFO) under most environmental conditions suggesting that the environmental partitioning of PFOA will be dominated by the anionic form. As the free acid readily dissociates to perfluorooctanoate (PFO), models for global fate, transport/partitioning and bioaccumulation may not have considered the pH of the environmental compartment in conjunction with the pK_a values of the compound. Thus, the distribution ratios among air, water and other compartments may not have been calculated correctly, and the speciation of PFOA/PFO may not have been considered appropriately.

The physical and chemical properties of APFO are summarized in Table 3.

Table 3. Physical/chemical Properties of PFOA Ammonium Salt (APFO)

PFOA and its salts are of anthropogenic origin with no known natural sources (Kissa 1994). Surveys of industry conducted in 2000 and 2004 under the authority of section 71 of CEPA 1999, were used to collect data on the Canadian manufacture, import and export of certain perfluoroalkyl and fluoroalkyl substances, their derivatives and polymers, including PFOA (Canada 2000b, Canada 2004).

Results from the 2000 survey indicated that PFOA and its salts were not manufactured in Canada. Approximately 600 000 kg of PFAs were imported into Canada between 1997 and 2000. Imports of PFOA and its salts were reported by one company. The import of PFOA and its salts (< 1000 kg) represented a very small proportion of the PFAs imported and consisted almost exclusively of the ammonium salt used in industrial applications. The volumes reported do not include quantities possibly imported in manufactured items (Environment Canada 2001). The reported uses for PFOA and its salts included use as a polymer, component of a formulation and other uses (i.e., batteries, coatings, lubricants) (Environment Canada 2001). Owing to its physicochemical properties, PFOA may be used as a replacement for perfluorooctane sulfonate (PFOS) in applications that had not previously used PFOA (US EPA 2002).

Industrial information was also obtained for the 2004 calendar year on the Canadian manufacture, import and export of perfluoroalkyl and fluoroalkyl (PFA/FA) substances, including PFCAs (Canada 2004). Information from this survey indicated no known manufacture of PFA/FA substances in Canada. The survey also indicated that the PFOA ammonium salt was imported into Canada in quantities ranging between 100 – 100 000 kg (Environment Canada 2005) under the North American Industry Classification System (NAICS) code “Chemical and Allied Product Wholesaler”.

PFOA has been synthesized by two main industrial processes: Simons electrochemical fluorination and telomerization (US EPA 2002). The first method relies on an electric current that is passed through a solution of anhydrous hydrogen fluoride containing, typically, an octanoic acid derivative. In this manner, all hydrogen atoms are replaced by fluorine, yielding 30–45% linear perfluorooctanoyl fluoride, along with a variable mixture of other isomers, homologues and by-products (US EPA 2002). The perfluorooctanoyl fluoride is then separated and hydrolyzed to yield PFOA. The salts of PFOA are prepared by the base neutralization of the parent acid using the appropriate metal-containing base. The second method, telomerization, involves the reaction of a molecule known as a telogen (e.g., pentafluoroethyl iodide) with two or more molecules called taxogens (e.g., tetrafluoroethylene). This process yields telomer iodides, which can be oxidized to produce a carboxylic acid--for instance, >99% pure PFOA (Prevedouros et al. 2006)--and reacted with ethylene to produce a further iodide. These iodides can then be reacted to form a multitude of functional materials. The telomerization process generally results in mixtures of compounds with even ranges of carbon numbers.

APFO is used primarily as a commercial polymerization aid in the manufacture of fluoropolymers such as polytetrafluoroethylene and polyvinylidene fluoride (US Government 2003; OECD 2006; Prevedouros et al. 2006), which are used in various sectors, including the automotive, electronics, construction and aerospace industries. Fluoropolymers are used in the manufacture of stain- and water-resistant coatings on textiles and carpet; hoses, cable and gaskets; non-stick coatings on cookware; and personal care products (US Government 2003). APFO is also used as a constituent in aqueous fluoropolymer dispersions, which are formulated into paints, photographic film additives and in the textile finishing industry (OECD 2006). Aqueous fire-fighting foams may also contain APFO as a component (OECD 2006; Prevedouros et al. 2006). Fluorochemicals that are potential PFOA precursors are used in the treatment of food packaging materials to enhance their properties as a barrier to moisture and grease (Begley et al. 2005). Thus, although APFO is typically not intended to remain in manufactured articles, trace amounts may be present as a contaminant or degradation product.

Releases to the environment may occur during manufacturing and processing operations and throughout the service life and subsequent disposal of articles containing PFOA. Potential point sources thus include direct releases from manufacturing or processing facilities. Indirect releases may result from the degradation or transformation of precursors in wastewater treatment plants (WWTPs) and landfills. Such precursors may include parent compounds or chemical products containing PFOA. Potential precursors include related fluorochemicals that are detectable in the atmosphere (e.g., 8:2 fluorotelomer alcohols [FTOH], which have eight fluorinated carbons and a two-carbon ethyl alcohol group) and can degrade or transform to PFOA through biotic or abiotic pathways.

Direct Releases

PFOA and its salts are not manufactured in Canada (Environment Canada 2001). There are no published data on direct releases to air, water or land from Canadian industrial facilities (Ellis et al. 2004b).

Muir and Scott (2003), Scott et al. (2003) and Boulanger et al. (2005) reported the presence of PFOA in sewage treatment plant (STP) effluents entering the Great Lakes. Measured PFOA concentrations in treated effluents from STPs in Thunder Bay and Sault Ste. Marie, Ontario, ranged from 7.9 to 24 ng/L (Scott et al. 2003). PFOA was also measured in a North Toronto STP effluent at 38 ng/L (Muir and Scott 2003). Crozier et al. (2005) measured PFOA in effluent waters (concentrations ranging from 7 – 55 ng/L) and biosolids (concentrations ranging from 0.7 – 0.9 ng/g) from Ontario sewage treatment plants. Crozier et al. (2005) also noted that PFOA was detected at 7 ng/L in one STP influent and then measured at 7 ng/L in the same STP’s effluent suggesting that PFOA may have formed during the sewage treatment plant processes.

Boulanger et al. (2005) conducted a lakewide mass budget analysis of eight perfluorooctane surfactants, includingPFOA, for Lake Ontario. Boulanger et al. (2005) cited and used a 3M Company 1999 study which analyzed the finished effluents from six wastewater treatment plants (WWTP) for three perfluorinated compounds, including PFOA. Four WWTPs were located in US cities with known sources of manufacture or industrial use of perfluorinated compounds, and two WWTPs were located in US cities with no known sources. All WWTPs had detectable concentrations of PFOA, ranging from 41.2 to 2420 ng/L. Boulanger et al. (2005) used a PFOA concentration in WWTP effluent of 549 ± 840 ng/L from this 3M Company 1999 study for the mass budget analysis. Boulanger et al. (2005) noted that there was a high degree of uncertainty associated with this concentration owing to the limited number of samples studied, and the study did not include actual WWTP effluent discharge to the Great Lakes system. The Boulanger et al. (2005) mass budget calculations showed the inflow from Lake Erie and wastewater discharges to be the major sources, whereas outflow through the St. Lawrence River was the dominant loss mechanism, indicating that there are unaccounted sources to Lake Ontario. Boulanger et al. (2005) suggested that the cleaning and care of surface-treated products by consumers and uses of perfluorooctane surfactants in industrial processes may lead to the presence of these substances in WWTP discharges. It was also suggested that discarded treated articles in landfills and the subsequent treatment of landfill leachates by municipal water treatment works may introduce significant amounts of perfluorooctane surfactants into the environment. Dinglasan-Panlilio and Mabury (2006) found that different fluorinated materials and/or products all contained free or unbound fluorinated alcohols. The authors suggested that the residual fluorinated alcohol contribution to the atmospheric load of telomer alcohols is significant and may be the dominant source, given that the release of these residual fluorinated alcohols may occur all along the supply chain from production through application to actual consumer use.

There are some Canadian data on releases of PFOA from landfills (Ikonomou 2006). PFOA was detected in the Arctic landfill sediment at 22–1083 ng/g and in Kamloops, British Columbia landfill sediments up to 186 ng/g. Landfill leachates from across Canada were also analyzed forPFOA.PFOA was detected in landfill leachates in Waterloo, Ontario (458 ng/L), Cambridge, Ontario (1144 ng/L), Moncton, New Brunswick (88 ng/L), Halifax, Nova Scotia (2040 ng/L), Charlottetown, Prince Edward Island (642 ng/L), Toronto, Ontario (880 ng/L), Kelowna, British Columbia (146 ng/L), and Calgary, Alberta (238 ng/L). PFOA has been detected in landfill leachates (91.3–516 ng/L) (Kallenborn et al. 2004) in other countries that, like Canada, do not manufacture perfluorochemicals.

Indirect Releases

Potential sources for the formation of PFOA, such as the degradation or transformation of precursors, could lead to indirect environmental releases and contribute to the total amount of PFOA found in the environment.

D’Eon et al. (2007) determined that PFOA can be formed from polyfluoroalkyl phosphate surfactants (PAPS) such as 8:2PAPS via cleavage of the phosphate ester linkage, releasing free 8:2 FTOH, with subsequent biotransformation to PFOA.

PAPS are approved by the U.S. EPA as an inert defoaming additive to pesticide formulations and as nonpolymeric fluorinated surfactants approved for application to food contact paper products (D’Eon et al. 2007).

Wallington et al. (2006) used a three-dimensional global atmospheric chemistry model (IMPACT) to indicate thatn-C₈F₁₇CH₂CH₂OH (i.e., 8:2 FTOH) degrades in the atmosphere to give PFOA and other PFCAs. Schenker et al. (2008) used a global-scale multispecies model (CliMoChem) to indicate that, until the year 2000, the contribution of atmospheric fluxes from perfluorooctanesulfonyl fluoride–based substances to the atmospheric deposition of PFOA in the Arctic was similar to the contribution from fluxes from FTOHs. Depending on the location and season, molar PFOA concentrations in the atmosphere are considered to be the correct order of magnitude to explain observed levels in Arctic biota (Wallington et al. 2006). The seasonal behaviour of PFOA is such that relatively high PFOA concentrations (>1.5 × 10³ molecules/cm³) extend throughout the Arctic during the Arctic summer, whereas winter PFOA concentrations are lower by an order of magnitude (Wallington et al. 2006).

The formation of PFOA through the thermolysis of fluoropolymers has been reported by Ellis et al. (2001, 2003). The results from these studies indicate the potential for this process to produce PFOA. However, Ellis et al. (2001, 2003) stated that this process is unlikely to release significant quantities of PFOA to the environment and would not contribute to its long-range transport. The onset of thermolysis of fluoropolymers occurs at 365°C (not an environmentally relevant temperature). These temperatures could be reached in industry and in household applications and, as such, the thermolysis of fluoropolymers could be considered a source of PFOA.

FTOHs have been shown to metabolize to PFOA in rats (Hagen et al. 1981). A ß-oxidation mechanism was proposed by Hagen et al. (1981) to account for the formation of PFOA. Dinglasan et al. (2004) showed aerobic degradation of 8:2 FTOH with an initial half-life of ~0.2 day per milligram of initial biomass protein followed by a second half-life of 0.8 day per milligram in a mixed microbial culture obtained from sediment and groundwater taken from a contaminated site, enriched on 1,2-dichloroethane and subsequently maintained using ethanol as the sole carbon source. This mixed culture was chosen because it was acclimated to degradation of chlorinated alkanes and alcohols and, therefore, considered to be active on fluorinated alcohols. The degradation of the alcohol occurred primarily through a mechanism leading to a telomer acid, which then underwent ß-oxidation to yield PFOA, accounting for 3% of the total mass of FTOH initially at day 81. However, because this study was limited to identification and quantification of known or predicted transformation products, potential unknown transformation products were not identified. FTOHs were capable of metabolizing to PFOA in municipal wastewater treatment sludge (Pace Analytical 2001). Liu et al. (2007b) showed the microbial transformation of 8:2 FTOH to PFOA in clay soil and two pure soil bacterial cultures (Pseudomonas species).

Wang et al. (2005) conducted aerobic biodegradation studies of¹⁴C-labelled 8:2 FTOH in diluted activated sludge from a WWTP. Three transformation products were identified: 8:2 saturated acids, 8:2 unsaturated acids and PFOA, representing 27%, 6.0% and 2.1%, respectively, of the initial ¹⁴C mass after 28 days. Results suggested that perfluorinated acid metabolites such as PFOA account for only a very small portion of the transformation products observed over the time frame considered (Wang et al. 2005). Wang et al. (2005) also suggested that the biological fate of 8:2 FTOH is determined by multiple degradation pathways, with neither ß-oxidation nor any other enzyme-catalyzed reaction as a single dominant mechanism. A study by Dinglasan et al. (2005) showed that the oxidation of the 8:2 FTOH to the telomer acid occurred via the transient telomer aldehyde. The telomer acid was then further transformed via a ß-oxidation mechanism, leading to the unsaturated acid and PFOA. However, a complete mass balance was not achieved, and the authors attributed this to binding of metabolites to biomass and other biological macromolecules, unaccounted metabolites, uptake of intermediates (formation of covalent linkages) or alternative degradation pathways (Dinglasan et al. 2005).

FTOHs may be released from polymeric materials or chemicals that incorporate FTOHs, or residual amounts of FTOHs that failed to be covalently linked to polymers or chemicals during production may be released. FTOHs are used in fire-fighting foams, personal care and cleaning products, and oil, stain, grease and water repellent coatings on carpet, textiles, leather and paper (US EPA 2006a). FTOHs are also used in the manufacture of a wide range of products, such as paints, adhesives, waxes, polishes, metals, electronics and caulks. During the years 2000–2002, an estimated 5 × 10⁶ kg of these compounds per year were produced worldwide, 40% of which was in North America (Dinglasan et al. 2004). Fluorotelomer-based raw materials and products are manufactured by a series of steps, beginning with Telomer A. Global Telomer A production between 2000 and 2002 was between 5000 and 6000 tonnes per year (Prevedouros et al. 2006).

Measured vapour pressures of FTOHs range from 140 to 990 Pa. The calculated dimensionless Henry’s Law constants for this class of compounds (e.g., 270 at 25°C for 8:2 FTOH) using the limited data available for water solubility and vapour pressure reveal the propensity of these compounds to partition to air (Dinglasan et al. 2004). Ellis et al. (2004a) showed the potential for FTOHs to react in the atmosphere with hydroxyl radicals to yield PFOA. Smog chamber studies indicate that FTOHs can degrade in the atmosphere to yield a homologous series of PFCAs (Ellis et al. 2004a). It is believed that oxidation of FTOHs in the atmosphere is initiated by reaction with hydroxyl radicals (Dinglasan et al. 2004; Ellis et al. 2004a). It was shown that perfluorooctyl sulphonamides react with hydroxyl radicals to yield PFOA (Hatfield et al. 2002). Although these experiments were conducted at environmentally unrealistically high concentrations of hydroxyl radicals and the results were qualitative rather than quantitative, they do show the potential for related products to react atmospherically to producePFOA.

Stock et al. (2004, 2007) showed that there is a significant concentration of FTOHs present and widely disseminated in the North American atmosphere. A recent air sampling campaign detected FTOHs at tropospheric concentrations typically ranging from 17 to 135 pg/m³, with urban locations having higher concentrations than rural areas (Martin et al. 2002; Stock et al. 2004). Loewen et al. (2008) studied atmospheric concentrations of FTOHs and lake water concentrations over an altitudinal transect in western Canada. Lake water samples were collected at Cedar Lake (a small lake near Golden, British Columbia), at Bow Lake in Banff National Park (Banff, Alberta) and at another unnamed small lake in Banff National Park (Banff, Alberta). Passive air samplers were deployed on altitudinal transects (800–2740 above sea level) from Golden, British Columbia, to Banff National Park. Loewen et al. (2008) noted that the amount of 8:2 and 10:2 FTOHs (< 2.0 ng/sampler) increased with increasing altitude. Lake water concentrations of PFOA along the elevation transect were below 1 ng/L. No clear trend was evident between altitude and PFOA concentrations. Ellis et al. (2004a) and Wallington et al. (2006) indicated that telomer alcohols may be responsible in part for the presence of PFCAs in the Arctic and other non-urban areas where concentrations of peroxy radicals far exceed those of nitrogen oxides. It was noted that since the reaction of 8:2 FTOH with nitric oxide competes with the reaction that forms PFOA, the formation of PFOA should decrease with increasing nitrogen oxide concentrations. The production of PFOA is therefore suppressed in source regions that typically have nitrogen oxide concentrations of 100 parts per trillion (ppt) or greater.

In conjunction with the atmospheric measurements of the alcohols made by Martin et al. (2002) and Stock et al. (2004) and with the profile of the linear to branched isomers observed in Canadian Arctic samples (De Silva and Mabury 2004), Ellis et al. (2004a) concluded that telomer alcohols were a plausible source for some PFOA in remote regions. This conclusion is supported by Stock et al. (2007), who measured FTOHs in air at Cornwallis Island, Nunavut, in 2004. Mean concentrations of FTOHs ranged from 2.8 to 14 pg/m³. This conclusion is also supported by observations of PFOA in US rainwater (Scott et al. 2003, 2006b). These results indicate that FTOHs are widely disseminated in the troposphere and are capable of long-range atmospheric transport. In addition, De Silva et al. (2009) detected branchedPFOA isomers in Arctic and Lake Ontario sediment and surface water, Lake Ontario biota and humans; however, branched PFOA isomers were not detected in ringed seals and polar bears above the detection limits (3.6 ng/g).

Environmental modelling of PFOA and its salts cannot be conducted using widely accepted models, such as the Level III fugacity model, as they cannot be applied to ionic surfactants. The high water solubility of PFOA, coupled with the negligible volatility of ionized species, suggests that all PFOA species will partition primarily to the aquatic environment. Based upon experimental results, PFOA may have some ability to re-enter the gas phase from water (US EPA 2002). However, studies have shown that if this process occurs, then it occurs to a negligible extent around pH 8.5 (Oakes et al. 2004). Once in the aqueous phase, PFOA may partition to sediments, as is evident by its measurement in this medium (Giesy and Newsted 2001; Stock et al. 2007). However, a comparison of concentrations observed in the aqueous phase and in sediments suggests that sediments are unlikely to be a major sink for PFOA. This is supported by observations made by Oakes et al. (2004) and Masunaga and Odaka (2005). The adsorption and desorption properties of APFO were investigated in one activated sludge sample and four soil samples (Dekleva 2003). Dekleva (2003) found that the average adsorption of APFO ranged from 40.8% to 81.8%. Adsorption coefficient (K_d) values ranged from 0.41 to 36.8 mL/g. Organic carbon adsorption coefficient (K_oc) values ranged from 48.8 to 229 mL/g, and organic matter adsorption coefficient (K_om) values ranged from 28.4 to 133 mL/g. These values indicate that PFOA is more likely to sorb to organic carbon in soils than to other soil solids. Moodyand Field (1999) suggested that since PFOA could be measured in groundwater in areas where PFOA is no longer used, a fraction of the compound must be bound to soil and slowly released to water. However, it is recognized that the presence in groundwater may simply reflect the slow migration of PFOA rather than soil binding. It is unlikely that terrestrially deposited PFOA will undergo any long-range transport (Franklin 2002).

The information below was considered in evaluating whether PFOA meets the criteria for persistence and bioaccumulation as defined under the Persistence and Bioaccumulation Regulations of CEPA 1999 (Canada 2000a). Persistence criteria are half-lives of greater than or equal to 2, 182, 365 or 182 days for air, water, sediment and soil, respectively. Bioaccumulation criteria are bioaccumulation factors (BAFs) or bioconcentration factors (BCFs) of greater than or equal to 5000 or a log octanol–water partition coefficient (K_ow) of greater than or equal to 5.0.

Persistence

Available data indicate that PFOA does not significantly photodegrade under relevant environmental conditions (Todd 1979; Nubbeet al. 1995; Scrano et al. 1999; Hatfield 2001; Hori et al. 2004, 2005, 2008), hydrolyze (Ellis et al. 2004b) nor undergo abiotic or biotic degradation (Reiner 1978; 3M Company 1979, 1980b, 1985b; Pace Analytical 1997; Oakeset al.2004; Moriwaki et al. 2005; Cheng et al. 2008). In summary, PFOA has been shown to be persistent (Table 4), with studies indicating no abiotic or biotic degradation in the environment under relevant environmental conditions.

Table 4. Summary of Persistence Data

Half-life values for PFOA and its salts are estimated and remain highly speculative owing to the short study periods. The atmospheric lifetime of PFOA has been predicted to be 130 days (Hurley et al. 2004). Franklin (2002) calculated an atmospheric lifetime of PFOA to be in the order of days when it was emitted from a ground source, and therefore likely not subject to long-range transport. However, if PFOA is produced from an atmospheric source (i.e., via precursors) and if the major loss mechanism is wet or dry deposition, then it may have a lifetime of 20–30 days before deposition (Ellis et al. 2004b). This would be sufficient time to allow transport over many thousands of kilometres, implying a long-range transport mechanism. Thepresence of PFOA in the Canadian Arctic also may provide evidence for the long-range transport of either PFOA (e.g., via ocean currents) (Caliebe et al. 2004; Yamashita et al. 2005) or volatile precursors to PFOA through the atmosphere (Stock et al. 2007). A suggested hypothesis for the presence of PFOA in biota in remote regions is that a precursor (e.g., FTOHs) is emitted to the atmosphere and ultimately degrades to yield PFOA through biotic and abiotic degradation. Ellis et al. (2004a) showed that the atmospheric lifetime of short-chain FTOHs, as determined by their reaction with hydroxyl radicals, was approximately 20 days. Piekarz et al. (2007) estimated that atmospheric residence times of 6:2 FTOH, 8:2 FTOH and 10:2 FTOH were 50, 80 and 70 days, respectively.

Shoeib et al. (2006) collected air samples during a crossing of the North Atlantic and the Canadian Arctic Archipelago in July 2005 to investigate concentrations of FTOHs. The highest concentrations of FTOHs were for 8:2 FTOH (5.8–26 pg/m³), followed by 10:2 FTOH (1.9–17 pg/m³) and then 6:2 FTOH (ND–6.0 pg/m³). Ju et al. (2008) measured PFOA in the sea microlayer (air/water interface) and subsurface seawater near the Dalian coastal waters in China. PFOA concentrations in the sea microlayer ranged from 0.26 to 1.19 ng/L. In the subsurface water, PFOA concentrations ranged from 0.17 to 0.67 ng/L. The detection of PFOA in oceanic waters suggests another potential mechanism for its long-range transport to remote locations such as the Canadian Arctic.

Mass calculations for marine transfer of PFO to the Arctic resulted in a flux between 2 and 12 tonnes per year (Prevedouros et al. 2006). Armitage et al. (2006) estimated a net PFO flux of between 8 and 23 tonnes per year to the Arctic. However, as stated previously, based on the modelling assumptions made regarding the pK_a, pH and PFOA/PFO speciation, partitioning processes related to the neutral acid may be ignored, and this may result in an underestimation of the concentrations of PFOA (Burns et al. 2008; Goss 2008). PFOA was measured in polar ice caps from three areas in the High Arctic (Melville ice cap, Northwest Territories; Agassiz ice cap, Nunavut; and Devon ice cap, Nunavut) (Young et al. 2007). PFOA concentrations ranged from 0.012 to 0.147 ng/L, suggesting that contamination may be a result of atmospheric input. Between 1996 and 2005, there was no significant trend of PFOA concentrations (regression analysis, p = 0.140) (Young et al. 2007). Fluxes were calculated using the density corrected concentration, multiplied by the yearly accumulation. Fluxes calculated to each of the ice caps were multiplied by the area of the Arctic to yield a flux of PFOA to the area north of 65°N. These fluxes are estimates and may not be representative of actual deposition in this region due to wide variations in precipitation rates. PFOA showed a flux ranging between 114–587 kg/year in 2005 (Young et al. 2007).

Bioaccumulation

The log K_ow has been calculated or modelled for PFOA and its various salts. A modelled value has been determined to be 5 ± 0.5 (Jasinski et al. 2009). However, the K_ow is a problematic parameter for ionized surfactants because of their tendency to aggregate at the interface of a liquid–liquid system. Therefore, the K_ow should not be used to model bioaccumulation of PFAs. Instead, more emphasis should be placed on the results of experimental studies. In most PFOA bioaccumulation studies examined, BCFs and BAFs are below the “5000” bioaccumulation criterion stipulated in the Persistence and Bioaccumulation Regulations of CEPA 1999 (Canada 2000a). However, the criteria upon which these regulations are based were developed over 10 years ago, as outlined in the 1995 Toxic Substances Management Policy (Canada 1995a) and the Toxic Substances Management Policy: Persistence and Bioaccumulation Criteria (Canada 1995b). The criteria focused specifically on aquatic organisms (fish) and on neutral organic compounds, which, unlike perfluorinated substances, preferentially partition to lipids. PFOA is an ionized substance that primarily binds to proteins (Han et al. 2004) and preferentially partitions to liver and blood and kidney tissues. The numeric criteria for bioaccumulation, outlined in the Persistence and Bioaccumulation Regulations of CEPA 1999, are based on bioaccumulation data for aquatic species (fish) only and for substances that preferentially partition to lipids. As a result, the criteria do not account for the bioaccumulation of PFOA that is preferentially partitioning in the proteins of liver, blood and kidney in terrestrial and marine mammals. Notably, the Toxic Substances Management Policy (Canada 1995a) states that expert opinion and a weight of evidence approach are also very important when considering how to interpret and apply the criteria.

PFOA is absorbed in juvenile rainbow trout (Oncorhynchus mykiss), with a BCF of 4.0 (Martin et al. 2004b). Dietary exposure to PFOA in the same species did not result in PFOA bioaccumulation (BAF = 0.038) (Martin et al. 2004b). Carp (Cyprinus carpio) were exposed to two concentrations of PFOA (5 and 50 µg/L) for 28 days in a flow-through system (Kurume Laboratory 2001). The chemical was prepared for exposure by using a dispersant made up of hydrogenated castor oil and mixed in acetone. Water quality was monitored daily for the duration of the experiment. BCFs ranged from 3.1 at the low concentration to <5.1–9.1 at the high concentration, indicating low bioaccumulation (Kurume Laboratory 2001). Although experiments with fish and other aquatic species provide evidence that PFOA is not highly bioaccumulative, these results should not be extrapolated to other animals. Fish gills may provide an additional mode of elimination and uptake that air-breathing organisms such as birds, terrestrial organisms and marine mammals do not possess (Kelly et al. 2004). The high water solubility of PFOA causes its tendency to escape from the gills into water to be relatively high, whereas the tendency of PFOA to escape to air across the alveolar membrane of the lung would be relatively low because of its low vapour pressure and negative charge. For example, the half-life of PFOA in fish was 3 days (Martin et al. 2003a), whereas the half-life in male rats was 11 days (Ylinen and Auriola 1990); in humans, it was 4.37 years (KudoandKawashima 2003) and 3.8 years (Olsen et al. 2007).

Two species of wild turtles (red-eared slider, Trachemys scripta elegans; and Reeves’ turtle, Chinemys reevesii) were examined for BCFs. In Japan, these turtles occupy the highest trophic level of the food chain in the river ecological system and have small territories (Morikawa et al. 2006). PFOA concentrations in surface water ranged from 16.7 to 87 100 ng/L, and PFOA was observed in almost all serum samples (91 of 94), where concentrations ranged from <200 to 870 000 ng/L. The calculated BCFs ranged from 0.8 to 15.8 (Morikawa et al. 2006). The authors noted that the reported BCFsdecreased as PFOA concentrations in surface water increased, suggesting that the absorption of PFOA from the gut might be a saturable process. However, it should be noted that these BCFs are actually BAFs, as the wild turtles’ exposure to PFOA was probably not limited to surface water only.

Kannan et al. (2005a) measured PFOA concentrations in a benthic food chain of the Great Lakes and found that despite relatively high concentrations of PFOA in water, it was not detected in invertebrates or fish. Preliminary (unpublished) results from one study showed a biomagnification factor (BMF) of 8 from ringed seal (Pusa hispida) liver to polar bear (Ursus maritimus) liver (Butt and Smithwick 2004). In the pelagic aquatic food web of Lake Ontario, PFOA concentrations did not increase with increasing trophic level (as determined by stable isotopes of nitrogen) (Martin et al. 2004b). In fact, PFOA concentrations were lower in the top-predator lake trout (Salvelinus namaycush) than in the invertebrate opossum shrimp (Mysis relicta) (Table 5). For example, a trophic magnification factor (TMF) for PFOA was calculated as 0.37 for the slimy sculpin (Cottus cognatus)–burrowing amphipod (Diporeia hoyi)relationship; for the upper end of the food web (Mysis relicta–alewife [Alosa pseudoharengus]–rainbow smelt [Osmerus mordax]–lake trout), the calculated TMF was 0.58 (Martin et al. 2004b). The different ends of the aquatic food web could indicate differences in benthic and pelagic natures of the relationships. TMFs less than 1 indicate no biomagnification. In another study, the trophic level–corrected BMFs were calculated for several Arctic biota (Tomy et al. 2004). Tomy et al. (2004) determined TMFs for the entire food web based on the relationship between d¹⁵ N and contaminant concentration. Trophic level was determined relative to the clam, which was assumed to have a trophic level of 2 (i.e., primary herbivore). For each individual sample of zooplankton, fish and marine mammal, trophic level was determined using the following relationship:

TL consumer = 2 + (d¹⁵ N consumer - d¹⁵ N clam) / 3.8

where TL consumer is the trophic level of the organism and 3.8 is the isotopic enrichment factor. The second method determined biomagnification factors (BMF_TL) for individual species corrected for trophic level:

BMF_TL = [predator] / [prey] / (TL predator/TL prey)

where [predator] and [prey] are the wet weight (ww) concentrations of analyte in the predator and prey species, respectively, and TL is the trophic level based on d¹⁵ N for the predator and prey. The resulting BMFs for PFOA were often greater than 1, suggesting a potential for PFOA to biomagnify for the clam (Mya truncata; Serripes groenlandica)–walrus (Odobenus rosmarus), Arctic cod (Boreogadus saida)–narwhal (Monodon monoceros) and cod–beluga (Delphinapterus leucas) food chain (Table 5). These results may reflect food web differences and perhaps not the bioaccumulation potential for PFOA (e.g., one food web had fish as a top predator and the other had a mammal).

Martin et al.(2004a) found that polar bears, which occupy the highest trophic level in the Canadian Arctic, have higher levels of PFOA than all other Arctic organisms examined. Butt et al. (2008) determined regionally-based ringed seal/polar bear BMF values for PFOA that ranged from 45 – 125. These regionally based BMFs were calculated by grouping ringed seal populations to corresponding similarly located polar bear populations in the Canadian Arctic. Tomy et al. (2009) determined PFOA TMFs for a marine food web in the western Canadian Arctic (Hendrickson Island and Holman Island) comprising of the Beaufort Sea beluga whale (Delphinapterus leucas), ringed seal (Phoca hispida), Arctic cod (Boreogadus saida), Pacific herring (Clupea pallasi), Arctic cisco (Coregonus autumnalis), a pelagic amphipod (Themisto libellula), and a Arctic copepod (Calanus hyperboreus). TMFs ranged from 0.1 (ringed seal/Arctic cod) to 2.2 (Arctic cod/Calanus hyperboreus).

Houde et al. (2005) reported BMFs of 1.8–13 using whole-prey homogenates and whole-body bottlenose dolphin (Tursiops truncatus) conjugate base concentrations in the bottlenose dolphin food web at Charleston, South Carolina. Kelly et al. (2009) found a TMF of 3.28 over the Canadian Arctic (Hudson Bay region) marine food web (macroalgae, bivalves, fish, seaduck and beluga whale). van den Heuvel-Greve et al. (2009) found BMFs of 3.8 and 23 for the benthic and the pelagic food webs with Harbour seals (Phoca vitulina) as the apex predator, respectively. BMFs for other species in the Westerschelde, Netherlands estuary ranged from 0.03 to 31.

Table 5. Summary of Bioaccumulation Data

Bioaccumulation (BCFs, BAFs, BMFs) may indicate either direct toxicity in organisms that have accumulated PFOA or indirect toxicity in organisms that consume prey containing PFOA (via food chain transfer). The minimum concentration of a substance in an organism that will cause an adverse effect (the critical body burden) is used to determine the potential to cause direct toxicity. From a physiological perspective, it is the concentration of a substance at the site of toxic action within the organism that determines whether a response is observed, regardless of the external concentration. In the case of PFOA, the site of toxic action is often considered to be the liver. However, when the potential for toxicity in consumer organisms is being determined, it is the concentration in the whole body of a prey item that is of interest, since the prey is often completely consumed by the predator (including individual tissues and organs, such as the liver and blood). Conder et al. (2008) noted that the concentration of perfluorinated acids on a whole body mass basis has been estimated to be 2–10 times lower than the concentrations of perfluorinated acids in blood and liver of trout. Although, the bioaccumulation potential of PFOA may be low in fish, the presence of detectable concentrations in higher trophic levels (polar bear, caribou, walrus) has generated concerns regarding the biomagnification potential of PFCAs, including PFOA, in food webs (Conder et al. 2008). Since perfluorinated substances partition to liver and blood, most field measurements for these substances have been performed on those individual organs and tissues. This is especially true for organisms at the higher trophic levels (e.g., polar bear), where whole-body analysis is not feasible for ethical reasons and due to the challenging logistics with respect to sampling and laboratory constraints. While it is feasible to measure whole-body BAFs on smaller species at lower trophic levels, the lower trophic status of the organism means that the estimated overall BAFs for perfluorinated substances may be underestimated. Thus, from a toxicological perspective, BCFs, BAFs and BMFs based on concentrations in individual organs, such as the liver, may be more relevant when the potential for direct organ-specific toxicity (i.e., liver toxicity) is being predicted. BCFs and particularly BMFs based on concentrations in whole organisms may provide a useful measure of overall potential for transfer up the food chain. As such, it is considered that PFOA may not be bioaccumulative in aquatic species but may be considered to bioaccumulate and biomagnify in terrestrial and marine mammals.

Ecological Exposure Assessment

Air

Loewen et al. (2008) studied atmospheric concentrations and lake water concentrations of FTOHs over an altitudinal transect in western Canada. Lake water samples were collected at Cedar Lake (a small lake near Golden, British Columbia), at Bow Lake in Banff National Park (Banff, Alberta) and at another unnamed small lake in Banff National Park (Banff, Alberta). Passive air samplers were deployed on altitudinal transects (800–2740 above sea level) from Golden, British Columbia, to Banff National Park. Loewen et al. (2008) noted that the amount of 8:2 and 10:2 FTOHs (<2.0 ng/sampler) increased with increasing altitude. Lake water concentrations of PFOA along the elevation transect were below 0.001 µg/L. No clear trend was evident between altitude and PFOA concentrations.

Stock et al. (2007) took air samples on Cornwallis Island, Nunavut where mean values of total concentrations of FTOHs ranged from 2.8 (10:2 FTOH) and 14 pg/m³ (8:2 FTOH). PFOA was also measured at a mean concentration of 1.4 pg/m³.

Shoeib et al. (2006) took twenty high-volume air samples during a crossing of the North Atlantic and Canadian Archipelago in July 2005 (Gothenburg, Sweden to Barrow, Alaska via the North Atlantic and Canadian Archipelago). The highest concentrations (sum of gas- and particle-phases) of FTOHs were for 8:2 FTOH at 5.8 -26 pg/m³, followed by 10:2 FTOH at 1.9-17 pg/m³ and 6:2 FTOH at below detection to 6.0 pg/m³. For comparison purposes, Shoeib et al. (2006) also collected air samples at a semi-urban site in Toronto in March 2006 where the mean 8:2 FTOH concentration in Toronto was 41 pg/m³. Dreyer et al. (2009) conducted high volume air sampling in the Atlantic Ocean, the Southern Ocean and the Baltic Sea. PFOA was detected in the particle fraction with a maximum concentration of 6 pg/m³. 6:2 FTOH and 8:2 FTOH were dominant in the gas-phase fraction. The concentrations of 8:2 FTOH were between 1.8 and 130 pg/m³. The sum of all the FTOHs (4:2 FTOH, 6:2 FTOH, 8:2 FTOH, 10:2 FTOH and 12:2 FTOH) ranged between 0.3–47 pg/m³

Water

Moody et al. (2002) measured PFOA concentrations in surface waters in Etobicoke Creek, a tributary to Lake Ontario, after an accidental spill of aqueous fire-fighting foam. Background concentrations were also measured at an upstream site. Results indicated downstream PFOA concentrations up to 11.3 µg/Ldue to the spill. PFOA was also detectable at the upstream site at lower concentrations (ND–0.033 µg/L).

Boulanger et al. (2004) reported PFOA concentrations in Lake Erie and Lake Ontario waters. Samples were acquired at a depth of approximately 4 m at four locations in both Lake Erie and Lake Ontario. Sampling sites were selected to sample urban-influenced and remote locations, as well as to provide a sample from eastern, central and western portions of each lake. Results indicated concentrations of 0.021–0.047 µg/L in Lake Erie and 0.015–0.070 µg/L in Lake Ontario. However, the study did not clearly delineate which samples were from remote or urban-influenced areas. Muir and Scott (2003) measured PFOA concentrations in Lake Superior, Lake Huron and Lake Ontario (exact locations not provided). Results ranged from 0.0015 to 0.0018 µg/L in Lake Huron and from <0.000 01 to 0.0007 µg/L in Lake Superior. Lake Ontario, at depths of 70–213 m, had PFOA concentrations ranging from 0.0023 to 0.011 µg/L. The highest concentration of 0.011 µg/Lwas sampled at a depth of 213 m. In a study by Furdui et al. (2005), measurements of PFOA were attempted in Lake Ontario, Lake Erie, Lake Huron and the North Channel (along the north shore of Lake Huron) using a method eliminating the extraction/concentration steps. With this method, PFOA was measured only in Lake Ontario and Lake Erie, with concentrations ranging from 0.002 to 0.007 µg/L. Scott et al. (2003, 2006b) showed the presence of PFOA in the tributaries of Lake Ontario and Lake Erie. The PFOA concentrations in six Lake Ontario tributaries (Welland Canal, Trent River, Black River, Don River, Genessee River and Oswego River) ranged from 0.0015 to 0.025 µg/L. The four Lake Erie tributaries (Grand River, Stoney Creek, Sandusk Creek and Talbot Creek) had a concentration range of 0.0016–0.0093 µg/L. Tributary water from urban areas on Lake Ontario had maximum PFOA concentrations of 0.02 µg/L (Myers et al. 2009)

In 2001, PFOA was measured in precipitation in three remote areas (Turkey Lakes, Ontario; Kejimkujik, Nova Scotia; and Chapais, Quebec), with concentrations ranging from <0.0005 – 0.0031 µg/L (Scott et al. 2006b). Rainwater samples collected in Winnipeg, Manitoba, did not detect PFOA (method detection limit [MDL] of 0.0072 µg/L) (Loewen et al. 2005). Loewen et al. (2005) suggested that this may have been due to insufficient atmospheric concentrations of PFOA and a relatively high MDL. Scott et al. (2006a) measured PFOA in precipitation across Canada from 2002 to 2004. In 2002, Kejimkijik (remote site) had PFOA concentrations ranging from <0.0001 to 0.0031 µg/L. In 2002, Algoma, Ontario (remote site), had PFOA concentrations ranging from <0.0001 to 0.0061 µg/L. In 2003–2004, two urban sites in Ontario (Egbert and north Toronto) had PFOA concentrations ranging from 0.0007 to 0.0111 µg/L. In 2002, Saturna Island, British Columbia (rural site), had PFOA concentrations ranging from <0.0001 to 0.002 µg/L.

Stock et al. (2007) measured lake water samples from three Arctic lakes (Amituk, Resolute, and Char) on Cornwallis Island, Nunavut, in 2003–2005. Concentrations of PFOA ranged from 0.0009 to 0.014 µg/L. Ahrens et al. (2009) measured PFOA along the longitudinal gradient from Las Palmas (Spain) to St. John’s, Newfoundland, and along the latitudinal gradient from the Bay of Biscay to the South Atlantic Ocean in the spring and fall of 2007. PFOA was not detected above the MDL of 0.0012 µg/L in the particulate phase or the deep water samples at 200 m and 3800 m. PFOA concentrations ranged from 0.000 004 0 to 0.000 229 µg/L at 11 m, 2 m and directly at the surface. In addition, Ahrens et al. (2009) noted that the concentrations of PFOA and perfluorononanoic acid (PFNA) were positively correlated, indicating that the sources of both compounds are related. Del Vento et al. (2009) measured up to 0.000448 µg/L PFOA in seawater and 3.4 x 10^-5 – 0.002282 µg/L PFOA in snow from the Amundsen Gulf.

PFOA was also the major fluorinated contaminant detected in oceanic waters from the Pacific and Atlantic oceans and from several coastal seawaters from Asian sites (Japan, Hong Kong, China and Korea) (Yamashita et al. 2005). PFOA was detected at concentrations ranging from 0.000 001 5 to 0.000 192 µg/L, followed by perfluorooctane sulfonate (PFOS) at 0.000 001 1–0.0577 µg/L. Yamashita et al. (2005) also found that deep seawater samples collected at depths greater than 1000 m in the Pacific Ocean and the Sulu Sea contained trace levels (values not provided) of PFOA. PFOA was also observed in the North Sea (estuary of the river Elbe, German Bight, southern and eastern North Sea) at concentrations ranging from 0.003 to 0.02 µg/L(Caliebe et al. 2004). In the open sea, PFOA was detected at 0.0005 µg/L (Caliebe et al. 2004). Dissolved PFOA was also detected in Tokyo Bay, with concentrations ranging from 0.007 to 0.0182 µg/L(Masunaga and Odaka 2005). Concentrations of PFOA in brine and sea-ice were in the same range as the snow concentrations.

Sediment

Stock et al. (2007) measured PFOA in sediment core samples taken in three remote Canadian Arctic lakes (i.e., Resolute,Char, and Amituk) on Cornwallis Island, Nunavut in 2003. Core samples taken at Char Lake had PFOA concentrations up to 0.0017µg/g dry weight (g-dw). Core samples taken at Resolute Lake had PFOA concentrations ranging from 0.0023–0.00095 µg/g-dw. Suspended sediment samples were taken from Niagara-on-the-Lake in the Niagara River and measured for PFOA. PFOA concentrations increased slightly from <0.0001 to <0.0003 µg/g-dw from 1980 to 2002 (Lucaciu et al. 2005). Open lake sediment from Lake Ontario had maximum PFOA concentration of 0.0066 µg/g (Myers et al. 2009).

Soil, Groundwater and Vegetation

There have been no reported measurements of PFOA in soils or groundwater in Canada to date.

PFOA was measured in soils collected in Dalton, Georgia, USA, at levels ranging from 0.055 to 0.174 µg/g (Ellington et al. 2005). PFOA has been detected in groundwater from sites associated with military fire-fighting training activities in the United States (Florida, Nevada and Missouri), where aqueous film-forming foams have been used (Schultz et al. 2004). Concentrations ranged from ND to 6570 µg/L. The West Virginia Department of Environmental Protection.(2003) measured concentrations of PFOA in groundwater samples from drinking water wells where the concentrations ranged from ND (<0.01 µg/L) to 23.6 µg/L. Murakami et al. (2009) measured PFOA in groundwater and springwater samples from 0 to 33 m below ground in the Tokyo (Japan) metropolitan area from September to November 2006. The concentration of PFOA ranged from 0.000 47 to 0.06 µg/L.

Biota

Studies from Canada have found the conjugate base of PFOA in a wide variety of wildlife. Martin et al. (2004b) measured PFOA in the whole body of a wide variety of biotic species from Lake Ontario, including benthic and pelagic invertebrates and various fish species. The concentration range was 0.001–0.09 µg/g-ww, with the highest concentration being found in the benthic invertebrate Diporeia hoyi in Lake Ontario (Martin et al. 2004b). This data from Lake Ontario shows that the benthic invertebrate, Diporeia hoyi, had higher concentrations of PFOA than any other organism in Lake Ontario (i.e., 90 µg/kg) (including the top-predator lake trout), indicating that sediments may be a reservoir for PFOA in this system (Martin et al. 2004b). This does not necessarily mean that PFOA partitions strongly into aquatic sediments; it is more likely that PFOA precursors partition to the sediments and subsequently release PFOA upon bacterial or abiotic processing (Martin et al. 2004b; Stock et al. 2007).

Following a spill of fire-fighting foam in Etobicoke Creek, Moody et al. (2002) measured PFOA in the liver of common shiners (Notropis cornuta) at concentrations ranging from 0.006 to 0.091 µg/g-ww. Forty-six lake trout whole-body homogenates were analyzed in 2001 from all five of the Great Lakes (Furdui et al.2007). PFOA concentrations ranged from 0.0007 µg/g-ww (Lake Erie) to 0.0024 µg/g-ww (Lake Michigan) (Furdui et al. 2007).

Martin et al. (2004a) also showed the presence of PFOA in polar bear liver (0.0029–0.013 µg/g-ww), whereas PFOA concentrations in other Arctic animals were below the detection limit (i.e., <0.002 µg/g). In Nunavut, trace levels of PFOA were measured in the liver of Arctic char (Salvelinus alpinus) (ND), burbot (Lota lota) (ND–0.0265 µg/g-ww), caribou (Rangifer tarandus) (ND–0.0122 µg/g-ww), ringed seal (ND–0.0087 µg/g-ww) and walrus (ND–0.0058 µg/g-ww) (Ostertag et al. 2009). Powley et al. (2008) did not detect PFOA in samples of zooplankton (Calanis hyperboreus, Themisto libellula, Chaetognatha), Arctic cod (Boreogadus saida), the blubber, blood or liver of ringed seal (Phoca hispida) and the blubber, blood or liver of bearded seal (Erignathus barbatus) taken near Banks Island (eastern edge of the Beaufort Sea in the Northwest Territories). However, the sample sizes were small, ranging from 1 to 5. The limit of detection used in this study was 0.0002 µg/g.

Studies in the United States have found the conjugate base of PFOA in a wide variety of biotic samples, including fish, clams, oysters, birds, mink and otters. In general, concentration ranges were from below the limit of detection to 1.9345 µg/g-ww--higher than most concentrations measured in Canadian biota--found at a Guntersville, Alabama, outfall location in gar liver (Giesy and Newsted 2001). In the common cormorants (Phalacrocorax carbo), concentrations were observed up to 0.450 µg/g-ww (Kannan et al. 2002). However, it was noted by Kannan et al. (2002) that for this colony of cormorants, the highest value (0.450 µg/g-ww) appeared to qualify as an outlier, as the concentration was 4.5 times greater than the standard deviation of the mean. In benthic algae collected at the Raisin River, the St. Clair River and Calumet River, PFOA could not be detected (detection limit 0.2 ng/g-ww; Kannan et al. 2005a). PFOA was found to be one of the dominant perfluorinated compounds (in addition to PFOS) in plasma of immature loggerhead sea turtles (Caretta caretta) (0.000 493–0.008 14 µg/mL) and immature Kemp’s Ridley sea turtles (Lepidochelys kempii) (0.002 77–0.004 25 µg/mL) captured in offshore waters of South Carolina, Georgia and Florida (Keller et al. 2005). Since sea turtles in the pelagic juvenile stage feed in areas distant from continental influences, the detection of perfluorinated compounds in these turtles may indicate contamination of an ocean basin (Keller et al. 2005). Keller et al. (2005) also found that PFOA concentrations were not significantly different between species, sexes, ages or geographical locations. However, given that the captured turtles were juveniles, sex and age differences would not be expected.

Studies in Japan, China, Taiwan, Korea have found the conjugate base of PFOA in a wide variety of biota. Tseng et al (2006) found PFOA ranging in concentrations from 0.12 to 0.34 µg/g in oysters (Crassostrea gigas), Japanese seaperch (Lateolabrax japonicus), and tilapia (Oreochromis sp.). Nakayama et al. (2008) did not detect PFOA (limit of quantification was 0.005 µg/g-ww) in the livers of wild common cormorants (Phalacrocorax carbo) from Lake Biwa in Japan. Concentrations of PFOA were measured in egg yolks of the little egret (Egretta garzetta), little ringed plover (Charadrius dubius) and vinous-throated parrotbill (Paradoxornis webbiana) collected around Lake Shihwa, Korea (Yoo et al. 2008). PFOA concentrations ranged from <0.0008 to 0.0543 µg/g-ww. Wang et al. (2008) measured concentrations of PFOA in waterbird eggs (black-crowned night herons (Nycticorax nycticorax), great egrets (Ardea alba), and little egrets (Egretta garzetta) in South China where PFOA concentrations ranged from <0.000 001 to 0.000 952 µg/g-ww.

Male wild rats (Rattus norvegicus) collected in Japan (i.e., a WWTP, a port, two industrial areas, a seafood market/port, a marketplace, and two landfill sites) (Yeung et al. 2009a) had whole-blood PFOA concentrations from 0.000 06 to 0.006 57 µg/mL. PFOA was detected in the serum of captive giant panda (Ailuropoda melanoleuca) and red panda (Ailurus fulgens) in China (Dai et al. 2006). Serum concentrations ranged from 0.33 to 8.20 µg/Lfor the red panda and from 0.32 to 1.56 µg/Lfor the giant panda. PFOA was measured in serum of the Chinese Amur tiger (Panthera tigris altaica) (Li et al. 2008b) found in northeastern China, far eastern Russia and North Korea. PFOA was found at concentrations of 0.000 04–0.000 18 µg/mL. In another study, PFOA was analyzed in the serum of captive Bengal tigers (Panthera tigris tigris) and African lions (Panthera leo) from Harbin Wildlife Park,China (Li et al. 2008a). PFOA concentrations were found to be higher in the African lion than in the Bengal tiger, suggesting different exposure uptakes or metabolic capabilities between the two species. In the Bengal tigers, PFOA ranged in concentration from ND to 0.000 097 8 µg/mL and in the African lions, PFOA concentrations ranged from 0.000 286 to 0.001 04 µg/mL.

Holmström and Berger (2008) did not detect PFOA in the adult liver, adult kidney, adult muscle, chick liver and egg of the common guillemot (Uria aalge) from the island of Stora Karlsö in the Baltic Sea. Additionally, PFOA was not detected in herring collected 150 km from Stora Karlsö (herring comprises a large part of the common guillemot diet). PFOA was not detected in the liver of the northern fulmar (Fulmarus glacialis) along the coast of Svalbard and BjØrnØya in the Barents Sea (Norwegian Arctic) (Knudsen et al. 2007). Whole-blood concentrations of PFOA were measured in waterfowl from the Gulf of Gdansk, Baltic Sea--common scoter (Melanitta nigra); common eider (Somateria mollissima); red-throated loon (Gavia stellata); razorbill (Alca torda); and long-tailed duck (Clangula hyemalis). Concentrations ranged from 0.000 05 to 0.0018 µg/mL (Gulkowska et al. 2005). The study also measured whole-blood PFOA concentrations in cod, which ranged from 0.000 05 to 0.0007 µg/mL. PFOA was detected in beaver liver collected from Poland at concentrations of 0.000 28–0.000 29 µg/g-ww (Taniyasu et al. 2005).

PFOA was not detectable in fish, birds or marine mammals from Greenland and the Faroe Islands, except for polar bear liver (<0.012 µg/g-ww) and ringed seal liver (<0.012 µg/g-ww); however, these values were below the limit of quantification (Bossi et al. 2005). Livers were collected from 35 polar bears from two known subpopulations in northern and western Alaska (Southern Beaufort Sea subpopulation, from Icy Cape to east of Paulatuk in Canada; and Chukchi/Bering Sea subpopulation, near Russia and western Alaska) between 1993 and 2002 (Kannan et al. 2005b). Adult male polar bears from the Southern Beaufort Sea subpopulation had PFOA concentrations ranging from 0.0013 to 0.013 µg/g-ww. Adult male polar bears from the Chukchi/Bering Sea subpopulation had PFOA concentrations ranging from 0.001 to 0.0042 µg/g-ww (Kannan et al. 2005b). The authors also indicated that since lipid normalization of concentrations did not reduce the data variability within a population, the data were analyzed on a wet weight basis. The authors also found a lack of significant age, sex or subpopulation differences for PFOA.

PFOA was not observed in liver, blubber, muscle or spleen tissues in harbour seals (Phoca vitulina) from the Dutch Wadden Sea (detection limit was 0.062 µg/g-ww) (Van de Vijver et al. 2005). PFOA was measured in the liver and serum of the Baikal seal (Pusa sibirica) from Lake Baikal, eastern Siberia, Russia (Ishibashi et al. 2008b). In male and female Baikal seal liver, the concentration of PFOA ranged between <0.0015 and 0.0039 µg/g-ww. In male and female Baikal seal serum, PFOA concentrations ranged between <0.000 33 and 0.0019 µg/g-ww (Ishibashi et al. 2008b).

PFOA was detected in the plasma (0.0006–0.163 µg/g-ww) of bottlenose dolphin populations of Delaware Bay (Delaware), Charleston (South Carolina), Indian River Lagoon (Florida) and Bermuda (Houde et al. 2005). Significant age and location interactions were noted for PFOA concentrations in plasma. PFOA concentrations in plasma significantly decreased with age in dolphins from the Charleston and Indian River Lagoon areas. PFOA was also measured in plasma, milk and urine from wild bottlenose dolphins from Sarasota Bay, Florida (Houde et al. 2006). The concentration of PFOA in plasma ranged from 0.0018 to 0.0068 µg/g-ww, the concentration in milk was 0.0013 µg/g-ww and the concentration in urine was below the MDL (0.000 06 µg/g-ww). The concentration of PFOA significantly decreased with blubber thickness (a biological parameter related to body condition and contaminant storage). PFOA was measured in liver samples of Indo-Pacific humpback dolphins (Sousa chinensis) and of finless porpoises (Neophocaena phocaenoides) stranded in Hong Kong between 2003 and 2007 (Yeung et al. 2009b). PFOA concentrations in the humpback dolphins ranged from 0.000 243 to 0.008 32 µg/g-ww. PFOA was detected in finless porpoises at concentrations ranging from <0.000 25 to 0.000 859 µg/g-ww. PFOA was also detected in Franciscana dolphin (Pontoporia blainvillei) and Subantarctic fur seal (Arctocephalus tropicalis) collected from southern Brazil (Leonel et al. 2008). Concentrations were less than 0.0002 µg/g for both species.

Temporal and Geographical Trends

Temporal trends in PFOA concentrations were not found in archived common guillemot (Uria aalge) eggs in the Baltic Sea, Iceland, the Faroe Islands, Sweden and Norway, as PFOA was not detected (Holmström et al. 2005; Löfstrand et al. 2008). However, Verreault et al. (2007) examined freshly laid whole eggs of two herring gull colonies (RØst and HØrnØya) in northern Norway and found that PFOA concentrations increased significantly between 1983 and 1993 for the RØst colony but not for HØrnØya colony. There was also an increase post-1993 in both colonies. It was noted that the eggs from the RØst colony had significantly higher PFOA concentrations compared with the HØrnØya colony in 1993 and 2003. Temporal trends were not found in liver samples in thick-billed murres (Uria lomvia) and northern fulmars (Fulmaris glacialis) from Prince Leopold Island in the Canadian Arctic (Butt et al. 2007a). Temporal trends in PFOA concentrations were also not found in lake trout collected between 1979 and 2004 from Lake Ontario (Furdui et al. 2008) and in two Canadian Arctic ringed seal (Phoca hispida) populations sampled from 1992 to 2005 (Butt et al. 2007b).

However, temporal trends in PFOA concentrations were found in polar bears from Canada’s Baffin Island, which showed an increase in PFOA contamination in their livers from 1972 to 2002 (Smithwick et al. 2006). PFOA doubling time in liver tissue was calculated to be 7.3 ± 2.8 years for Baffin Island polar bears and 13.9 ± 14.2 years for Barrow, Alaska, polar bears. Smithwick et al. (2006) noted that the sex and age of polar bears were not significantly correlated with PFOA concentrations. Dietz et al. (2008) subsampled liver tissue of 128 subadult (3- to 5-year-old) polar bears (collected from central East Greenland) from 19 sampling years within the period 1984–2006 for perfluorinated compounds, including PFOA. The authors found annual increases of 2.3% for PFOA.

PFOA was measured in liver tissue from 80 adult female sea otters (Enhydra lutris) found freshly dead and beached along the California coast. Concentrations of PFOA in liver, ranging from < 5 x 10^-6 to 0.000147 µg/g-ww, increased from 1992 to 2002 for these adult female sea otters (Kannan et al. 2006) However, PFOA was not found in the adult male sea otters (detection limit of 0.005 µg/g). The reason for this lack of detection is not known (Kannan et al. 2006).

Ecological Effects Assessment

Aquatic Organisms

The most sensitive pelagic organism was found to be the freshwater alga, Pseudokirchneriella subcapitata. The96-hour LOEC based on both growth rate and cell count was 2.0 mg/L (Ward et al. 1995b, d).

Several toxicity tests on the freshwater alga Pseudokirchneriella subcapitata were conducted (Elnabarawy 1981; Ward et al. 1995b, d, 1996c, e, h; Boudreau 2002; Thompson et al. 2004), and the 96-hour EC₅₀ values were determined to range from 4.9 to >3330 mg/L based on growth rate and from 2.9 to 1980 mg/L based on cell count. The 96-hour no-observed-effect concentrations (NOECs) ranged from 1.0 to 500 mg/L based on growth rate and from 0.99 to 210 mg/L based on cell count. Similarly, the 96-hour LOECs ranged from 2.0 to 1000 mg/L based on growth rate and from 2.0 to 430 mg/L based on cell count. A 14-day EC₅₀ value for this freshwater alga was determined to be 43 mg/L (Elnabarawy, 1981). Toxicity studies with this freshwater alga were conducted using both commercial mixtures of the ammonium and tetrabutylammonium salts of PFOA and high-purity PFOA, which may account for the large ranges observed. Another toxicity test (Boudreau 2002) was conducted using high-purity PFOA and the freshwater alga Chlorella vulgaris. This study determined a 96-hour IC₅₀ value (based on growth rate) of 116 mg/L. This rate indicates that there may be little difference in the sensitivities of the two freshwater algal species. Liu et al. (2008) used flow cytometric measurements to investigate the effects of PFOA on the membrane systems of the freshwater alga, Scenedesmus obliquus. PFOA did not inhibit algal growth at the maximum test concentration of 0.000 002 M (0.83 mg/L). However, the mitochondrial membrane potential was affected, and exposure to PFOA at a concentration of between 0.000 001 (0.41 mg/L) and 0.000 002 M (0.83 mg/L) caused an increase in the permeability of the membrane. This suggests damage to the mitochondrial function and membrane permeability at a concentration that did not result in the inhibition of algal growth.

Six Microtox toxicity tests were conducted on the bacterium, Photobacterium phosphoreum, using commercial mixtures of the ammonium and tetrabutylammonium salts of PFOA (3M Company 1987a, 1990a, 1996a, b, c; Beach 1995a). The 30-minute EC₅₀ values (based on rate of bioluminescence) ranged from 260 to 3150 mg/L. It is unclear why this range is so large; however, it may be a result of the lack of characterization of the commercial mixtures and their impurities.

One toxicity test (Boudreau 2002), using high-purity PFOA and the aquatic macrophyte, Lemna gibba,determined a 7-day IC₅₀ value (based on growth rate) of 80 mg/L.

Several toxicity tests have been conducted on the water flea, Daphnia magna, using commercial mixtures of PFOA, the ammonium and tetrabutylammonium salts of PFOA and high-purity PFOA (3M Company 1982, 1984, 1987b; Ward and Boeri 1990; Ward et al. 1995e, 1996a, d, g; Boudreau 2002; CIT 2003). The 48-hour median lethal concentration (LC₅₀) values ranged from 77 to 1550 mg/L, whereas the 48-hour EC₅₀ values (based on immobilization) ranged from 34 to 1200 mg/L. It is unclear why this range is so large. It may be a result of the lack of characterization of the commercial mixtures and their impurities, or, as suggested by Ward et al. (1996a), it may be due to inconsistencies in diet. The 48-hour NOECs (based on immobilization) ranged from 13 to 730 mg/L, whereas 21-day NOECs (based on survival and/or reproduction or length of parent) ranged from 13 to 89 mg/L. One study determined the 21-day EC₅₀ value (based on reproductive capacity) to be 39.6 mg/L (CIT 2003). A single toxicity study (Boudreau 2002), conducted using high-purity PFOA and the water flea, Daphnia pulicaria, determined the 48-hour LC₅₀ value and 48-hour EC₅₀ value (based on immobilization) to be 277 mg/L and 204 mg/L, respectively. These results indicate that there may be little difference in the sensitivities in the two Daphnia species. Kim et al. (2009) conducted acute toxicity, reproduction and embryo development tests in Daphnia magna. PFOA showed a 48-hr EC₅₀ at 253.5 mg/L. The NOEC was 100 mg/L. PFOA caused reduction of fecundity at 10 mg/L and induced embryo lethality (arrested egg development) and neonate deformities (curved or unextended spines and undeveloped second antenna) at 125 mg/L.

Li (2008) conducted toxicity tests on the freshwater planarian (Dugesia japonica), freshwater snail (Physa acuta), water flea (Daphnia magna) and green neon shrimp (Neocaridina denticulata). The 96-hour LC₅₀ for the freshwater planarian ranged from 318 to 357 mg/L, and the NOEC was 150 mg/L; for the freshwater snail, the 96-hour LC₅₀ ranged from 635 to 711 mg/L, and the NOEC was 250 mg/L. For Daphnia magna,the 48-hour LC₅₀ ranged from 166 to 198 mg/L, and the NOEC was 125 mg/L; and for the green neon shrimp, the 96-hour LC₅₀ ranged from 418 to 494 mg/L, and the NOEC was 250 mg/L.

Several toxicity tests have been conducted on the fathead minnow using commercial mixtures of both PFOA and the ammonium and tetrabutylammonium salts of PFOA (3M Company 1977, 1985a, 1987c; Elnabarawy 1980; Wardet al. 1995a, c, 1996b, f, i, j). The 96-hour LC₅₀ values ranged from 70 to 2470 mg/L, and the 96-hour NOEC values (based on mortality) ranged from 110 to 830 mg/L. It is unclear why this range is so large; however, this may be a result of the lack of characterization of the commercial mixtures and their impurities. Studies were conducted to investigate the toxicity of the ammonium salt with bluegill (3M Company 1978a, b), and the 96-hour LC₅₀ values were >420 and 569 mg/L.

Two studies have also been conducted that examined the toxicity of the sodium salt of PFOA (high purity) on pelagic communities as a whole, using microcosms consisting of a community of zooplankton and mixed with large invertebrates for a study period of 35 days (Sanderson et al. 2003, 2004). Results of these studies determined both individual and community LOEC values to range from 10 to 70 mg/L. Another study has been conducted that examined the toxicity of the sodium salt of PFOA to the aquatic macrophytes Myriophyllum sibiricum and M. spicatum using 12 000-L outdoor microcosms over a period of 35 days (Hanson et al. 2005). The treatments applied were 0.3, 1, 30 and 100 mg/L. The endpoints monitored included growth, biomass, root number, primary root lengths and number of nodes. The results indicated that the two species of Myriophyllum were similar in their sensitivity to PFOA. The NOECs for Myriophyllum spp. were greater than or equal to 23.9 mg/L (Hanson et al. 2005).

Terrestrial Organisms

According to Tominaga et al. (2004), the soil-dwelling nematode Caenorhabditis elegans has been shown to be a suitable test organism, showing both lethal and sublethal endpoints, in the ecotoxicological assessments of liquid and soil media. Tominaga et al. (2004) examined acute lethal toxicity and multigenerational sublethal toxicity (fecundity and reproduction) using PFOA concentrations of 0, 0.01 mM (4.14 mg/L), 0.1 mM (41.4 mg/L), 0.5 mM (207 mg/L), 1.0 mM (414.07 mg/L) and 5.0 mM (2.1 mg/L) for 48 hours. All concentrations up to 0.1 mM (41.4 mg/L) showed no acute lethality until 48 hours. Acute lethality appeared at concentrations greater than 0.5 mM (207 mg/L) and did not depend on the incubation time. EC₅₀s were calculated for 1 hour (3.85 mM or 1590 mg/L), 2 hours (2.80 mM or 1160 mg/L), 3 hours (2.70 mM or 1120 mg/L), 4 hours (2.65 mM or 1100 mg/L), 24 hours (2.75 mM or 1140 mg/L) and 48 hours (2.35 mM or 973 mg/L) (Tominaga et al. 2004). In the multigeneration test, generation–response and concentration–response relationships were not observed for PFOA.

O’Brien et al. (2009) recently reported that linear PFOA that had been injected into the air cell of white leghorn chicken eggs had no effect on embryonic pipping success at concentrations up to 10 mg/g, and PFOA accumulated in the liver of these embryos to concentrations greater than the initial whole-egg concentration. It was concluded that, based on these white leghorn chicken results, current environmental concentrations of PFOA are unlikely to affect the hatching success of wild birds. Yeung et al. (2009c) exposed one-day old male chickens to a mixture of perflurooctane sulfonate (PFOS), perflurooctanoic acid (PFOA) and perfluorodecanoate (PFDA) at doses between 0.1 mg/kg body weight and 1.0 mg/kg body weight for three weeks. It was concluded that exposure to a mixture of PFOS/PFOA/PFDA at 1.0 mg/kg body weight had no adverse effect on juvenile chickens. The half-life for PFOA was 3.9 days at both doses.

Benthic Invertebrates

Toxicity tests of PFOA and its salts have been conducted with mixed-liquor activated sludge (obtained from the Metro WWTP in St. Paul, Minnesota). It should be noted, however, that the bacteria in activated sludge were selected for their ability to thrive on anthropogenic chemicals, and, as such, toxicity tests using them may underestimate toxicity. In total, five toxicity tests were conducted on mixed-liquor activated sludge using commercial mixtures of the ammonium and tetrabutylammonium salts of PFOA (3M Company 1980a, 1990b, 1996d; Beach 1995b). The 3-hour EC₅₀ values (based on respiration inhibition) ranged from >1000 to >3320 mg/L. In addition, the 7-minute NOEC (based on respiration inhibition) was determined to be 1000 mg/L (3M Company 1980a).

In a study by MacDonaldet al. (2004), the toxicity of high-purity PFOA to the aquatic midge Chironomus tentans was investigated. No toxicity was observed at any of the concentrations tested; as such, the 10-day NOEC value (based on survival and growth) was determined to be 100 mg/L.

Terrestrial Plants

Li (2008) conducted seed germination and root elongation toxicity tests on lettuce (Lactuca sativa), cucumber (Cucumis sativus) and pakchoi (Brassica rapa chinensis). PFOA had no effect on cucumber seed germination, with both LC₅₀ and NOEC values greater than 2000 mg/L. The LC₅₀ and NOEC values for lettuce seed germination were 1734 and 1000 mg/L, respectively. The LC₅₀ and NOEC values for pakchoi seed germination were 579 and 250 mg/L, respectively. The EC₅₀ for root elongation for the three species ranged from 263 to 1254 mg/L. PFOA almost completely inhibited lettuce and pakchoi root growth at or above 1000 mg/L. NOECs for root elongation for the three species ranged from <62.5 to 250 mg/L.

Stahl et al. (2009) studied the soil-to-plant carryover of a mixture of PFOA/PFOS (perflurooctane sulfonate) on spring wheat, oats, potatoes, maize, and perennial ryegrass. Concentrations ranged from 0.25 to 50 mg/kg of PFOA/PFOS as an aqueous solution. PFOA concentrations were higher than PFOS in all plants except for potatoes with uptake/storage more intensive in the vegetative portion than the storage organ. Visible abnormalities were noted at concentrations > 10 mg/kg. At 25 – 50 mg/kg PFOA/PFOS, necrosis was observed in both oats and potatoes, a yellowing of the ryegrass leaves, and diminished growth for spring wheat.

Endocrine Effects

Liu et al. (2007a) used freshwater male tilapia (Oreochromis niloticus) as the in vitro model to detect the induction of vitellogenin. Vitellogenin is an egg yolk precursor protein expressed in females of fish, amphibians, reptiles (including birds), insects and the platypus. In the presence of substances that affect endocrine function, males can also express the vitellogenin gene. Cultured male tilapia hepatocytes were exposed to PFOA, 4:2 FTOH, 6:2 FTOH and 8:2 FTOH for 48 hours. A dose-dependent induction of vitellogenin was observed in PFOA- and 6:2 FTOH–treated cells, whereas vitellogenin remained unchanged for 4:2 FTOH and 8:2 FTOH. The estimated 48-hour median effective concentration (EC₅₀) values were 2.9 × 10^-5 M (12 mg/L) for PFOA and 2.8 × 10^-5 M (12.9 mg/L) for 6:2 FTOH. In the time course study, vitellogenin induction took place at 48 hours (PFOA), 72 hours (4:2 FTOH), 12 hours (6:2 FTOH) and 72 hours (8:2 FTOH) and increased further after 96 hours of exposure. Co-exposure to a mixture of individual perfluorinated compounds and 17ß-estradiol for 48 hours significantly inhibited 17ß-estradiol-induced hepatocellular vitellogenin production in a dose-dependent manner, except for 4:2 FTOH. The estimated 48-hour median inhibitory concentration (IC₅₀) values were 5.1 × 10^-7 M(0.21 mg/L) for PFOA, 1.1 × 10^-6 M (0.51 mg/L) for 6:2 FTOH and 7.5 × 10^-7 M (0.35 mg/L) for 8:2 FTOH. In order to further investigate the estrogenic mechanism, the hepatocytes were co-exposed to a mixture of PFOA and 6:2 FTOH plus the known estrogen receptor inhibitor tamoxifen for 48 hours. The overall results demonstrated that PFOA and FTOHs have estrogenic activities and that exposure to a combination of 17ß-estradiol and PFOA or FTOHs produces anti-estrogenic effects. The results of the estrogen receptor inhibition assay further suggested that the estrogenic effect of PFOA and FTOHs may be mediated by the estrogen receptor pathway in primary cultured tilapia hepatocytes.

Wei et al. (2008a) assessed the effects of PFOA on male and female rare minnows (Gobiocypris rarus) at concentrations of 3, 10 and 30 mg/L for 28 days. Exposure to PFOA at 3 mg/L elicited moderate hepatocellular hypertrophy in the livers of both male and female fish. Male rare minnows exposed to PFOA at 10 mg/L showed eosinophilic hyaline droplets in the cytoplasm of the hepatocytes; female rare minnows displayed more eosinophilic hyaline droplets in the cytoplasm of the hepatocytes, hepatocellular hypertrophy and vacuolar degeneration. Rare minnows exposed to PFOA at 30 mg/L showed severe hepatic histopathological changes and disruption of mitochondrial functions. The inhibition of the thyroid hormone biosynthesis genes and the induction of estrogen-responsive genes may indicate a role in endocrine function. Wei et al. (2008b) further identified the potential protein biomarkers for PFOA exposure in the livers of the rare minnows at 3, 10 and 30 mg/L for 28 days, finding the abundance of 34 and 48 protein spots altered in males and females, respectively. These proteins were involved in intracellular fatty acid transport, oxidative stress, macromolecule catabolism, the cell cycle, maintenance of intracellular Ca²⁺ homeostasis and mitochondrial function. Wei et al. (2007) studied the in vivo effects of waterborne PFOA on the expression of hepatic estrogen-responsive genes, vitellogenin, and estrogen receptor and on the gonadal development in freshwater rare minnow (Gobiocypris rarus). The study showed mature females exposed to 3, 10, and 30 mg/L PFOA for 28 d had degenerating vitellogenic-stage oocytes (atresia) in the ovaries. In males exposed to 10 mg/L PFOA, primary

growth–stage oocytes (pre-vitellogenic oocytes) developed in some testes. The number of sperm and various stages of germ cells within the spermatogenic cycle in the 10 and 30 mg/L PFOA treatments were lower than those in control males. PFOA increased hepatic vitellogenin concentration and induced testis-ova gonads in mature male rare minnows at 10 and 30 mg/L for 28 days. Wei et al. (2007) showed that PFOA can disrupt the activity of estrogen by inducing hepatic estrogen-responsive genes in males, although the mechanism of development of testes-ova in rare minnows by PFOA exposure is not known.

Other Effects

There are several studies showing the potential of PFOA to affect endocrine function in wildlife. Stevenson et al. (2006) examined the toxicity of PFOA with respect to the multixenobiotic resistance mechanism in the marine mussel, Mytilus californianus. This mechanism acts as a cellular first line of defence against broad classes of xenobiotics exporting moderately hydrophobic chemicals from cells via adenosine triphosphate (ATP)–dependent, transmembrane transport proteins. The most studied transporter is the P-glycoprotein, which is a fragile defence mechanism and can be compromised by some xenobiotics. This increased sensitivity, referred to as chemosensitization, arises from the ability of the P-glycoprotein to recognize and bind to multiple xenobiotic substrates, resulting in the saturation of the binding capacity. Non-toxic substances can also be chemosensitizers and cause adverse effects on organisms by allowing normally excluded toxic substances to accumulate in the cell. Stevenson et al. (2006) found that PFOA at 50 µM (0.02 µg/L) significantly inhibited the P-glycoprotein in Mytilus californianus and thus is a chemosensitizer for that organism. The study also showed that this inhibition was reversible once the marine mussel was removed from contamination and placed in clean seawater.

Ishibashi et al. (2008b) showed that PFOA activates the mammalian peroxisome proliferator–activated receptor a (PPARa) in the livers of Baikal seals--the first reported identification of PPARa complementary deoxyribonucleic acid (DNA) in an aquatic wildlife species. PPAR is a member of the ligand-activated nuclear hormone receptor superfamily. PPARa plays a critical physiological role as a lipid sensor and a regulator of lipid metabolism. The lowest-observed-effect concentration (LOEC) for PFOA was 62.5 µM (25.9 mg/L).

The potential impact of exposure to perfluorinated compounds on liver lesions was investigated in East Greenland polar bears (Sonne et al. 2008). Parameters included mononuclear cell infiltrations, lipid granulomas, steatosis, Ito cells and bile duct hyperplasia/portal fibrosis. The population consisted of 28 females and 29 males harvested by local hunters between 1999 and 2002. Liver samples were analyzed for PFOS, perfluorononanoic acid, perfluoroundecanoic acid, perfluorodecanoic acid, perfluorotetradecanoic acid, PFOA, perfluorooctanesulfonamide, perfluorodecanoate and perfluorohexanesulfonate. In 23 cases, the concentration of PFOA was below the detection limit (0.0012 µg/g-ww). Sixty-five percent of the polar bears had total PFA concentrations above 1 µg/g-ww. In female bears, the total PFA concentration ranged from 0.256 to 2.77 µg/g-ww; in male bears, the total PFA concentration ranged from 0.114 to 3.052 µg/g-ww. All PFA compounds in the analysis were summed, so a direct cause–effect correlation with a particular perfluorinated compound, such as PFOA, cannot be determined. East Greenland polar bears are also contaminated with other substances, such as organochlorines (polychlorinated biphenyls [PCBs], dichlorodiphenyltrichloroethane [DDT]) and mercury, which may function as confounding synergistic co-factors in the development of the lesions. The authors concluded that the statistical analysis did not answer the question of whether chronic exposure to perfluorinated compounds is associated with liver lesions in polar bears; however, these lesions were similar to those produced by perfluorinated compounds under laboratory conditions (Sonne et al. 2008).

The effect of PFOA on immune function and clinical blood parameters has been examined in bottlenose dolphins and sea turtles from Florida, Georgia and South Carolina. It should be noted that a direct cause–effect relationship cannot be clearly established, as there may be other co-occurring contaminants. The results revealed that there may be increases in indicators of inflammation and immunity in bottlenose dolphin blood parameters in relation to PFOA, suggesting that PFOA may alter biomarkers of health in marine mammals (Peden-Adams et al. 2004a). Examples of biomarkers analyzed in bottlenose dolphins include absolute numbers of lymphocytes, serum triglyceride, serum total protein, serum albumin, serum cortisol, C-reactive protein, lysozyme activity and B-cell proliferation (Peden-Adams et al. 2004a). Serum triglyceride exhibited stronger relationships to PFOA in females than in males. Lipopolysaccharide-induced lymphocyte proliferation (B-cell proliferation) had positive but weak correlations with PFOA in male bottlenose dolphins, and a strong correlation was observed between PFOA and lysozyme activity (a measurement of innate immunity) in the same species. Low levels of PFAs may also alter biomarkers of health in loggerhead sea turtles (Peden-Adams et al. 2004b). Examples of biomarkers analyzed in loggerhead sea turtles include plasma total protein, plasma globulin, T-cell proliferation, plasma lysozyme activity and B-cell proliferation (Peden-Adams et al. 2004b). PFOA and 8:2 telomer acid have been detected in the urine of bottlenose dolphins from populations located off the coasts of Florida and South Carolina (Houde et al. 2005). This study also found that PFOA concentrations in plasma in dolphins from the Charleston, South Carolina, and Indian River Lagoon, Florida, areas decreased with age.

Characterization of Ecological Risk

The approach taken in this ecological screening assessment is to examine various supporting information and develop conclusions based on multiple lines of evidence approach such as persistence, exposure, trends, ecological risk, inherent toxicity, bioaccumulation and widespread occurrence in the environment. Endpoint organisms have been selected based on analysis of exposure pathways. For each endpoint organism, a conservative (reasonable worst-case) predicted environmental concentration (PEC) and a predicted no-effect concentration (PNEC) are determined. The PNEC is arrived at by selecting the lowest critical toxicity value (CTV) for the organism of interest and dividing it by an application factor appropriate for the data point. A risk quotient (PEC/PNEC) is calculated for each of the endpoint organisms in order to estimate potential ecological risk in Canada.

Wildlife

Butenhoff et al.(2002) conducted a 26-week cynomolgus monkey oral gavage study for which the lowest-observed-adverse-effect level (LOAEL) was 3 mg/kg body weight (kg-bw) per day for males for serum levels showing reversible liver effects and relative liver weight increases without histopathological effects. This study reported a mean liver concentration of PFOA of 15.8 µg/g at week 27 for the 3 mg/kg-bw per day treatment group (i.e., at the LOAEL). This value was selected as the CTV. This CTV was divided by an application factor of 100 to give a PNEC of 0.158 µg/g or 158 µg/kg. The application factor was used to account for laboratory to field extrapolation, intra- and interspecies variation and extrapolation from a LOAEL to a no-observed-adverse-effect level (NOAEL). The datum selected for the PEC is the highest liver PFOA concentration in polar bear--i.e., 13 µg/kg-ww, from Sanikiluaq, Nunavut, Canada (Martin et al. 2004a).

The risk quotient (PEC/PNEC) for Canadian mammalian wildlife is 0.08 (13/158). The risk quotient is less than 1, indicating low likelihood of risk from exposures at current concentrations in the environment.

Pelagic Organisms

Current Canadian surface water measurements include those of Etobicoke Creek (Toronto, Ontario), relating to a spill of aqueous fire-fighting foam (Moody et al. 2002), and water measurements in Lake Ontario (Boulanger et al. 2004). These data were selected for the PECs in Canada under three scenarios:

1. spill conditions (11.3 µg/L; downstream of spill over 150 days);
2. a creek in a densely populated region (0.033 µg/L; upstream of spill over 150 days); and
3. a lake in a densely populated region (highest background concentration 0.070 µg/L; Lake Ontario).

The spill condition is an extreme worst-case PEC. The upstream conditions in Etobicoke Creek may also be considered a worst-case PEC for general surface waters in Canada owing to the population density surrounding Etobicoke Creek, numerous WWTPs and storm sewer inputs along its length and the reasonably low natural flow of the creek, which causes minimal dilution of the anthropogenic inputs. Nonetheless, higher concentrations were measured in Lake Ontario. The maximum observed PFOA concentration in each scenario was selected as the PEC:

1. spill (PEC = 11.3 µg/L);
2. receiving creek (PEC = 0.033 µg/L); and
3. receiving lake (PEC = 0.070 µg/L).

Most of the available toxicity data are for freshwater pelagic organisms, given that PFOA is expected to partition primarily to the aquatic environment. The organism that was most sensitive to PFOA, as determined from single-species tests, was the freshwater alga Pseudokirchneriella subcapitata (96-hour LOEC for growth rate and cell count of 2.0 mg/L)(Ward et al. 1995b, 1995d).This 2.0 mg/L value is selected as the CTV for pelagic organisms. This CTV is divided by an application factor of 100 to account for laboratory to field extrapolation, potential effects caused by the presence of additional stressors, and intra- and interspecies variation, to give a PNEC of 0.02 mg/L or 20 µg/L. The risk quotients (PEC/PNEC) for pelagic organisms are as follows:

1. spill condition is 0.56 (11.3/20).
2. receiving creek conditions is 0.002 (0.033/20).
3. receiving lake conditions is 0.004 (0.07/20).

These risk quotients indicate low likelihood of risk to pelagic organisms from exposures at current concentrations in the environment.

Wei et al. (2007) showed that female rare minnows (Gobiocypris rarus) exposed to 3, 10, and 30 mg/L PFOA for 28 d had degenerating vitellogenic-stage oocytes in the ovaries. The value of 3 mg/L was selected as the CTV. This CTV is divided by an application factor of 100 to account for laboratory to field extrapolation, potential effects caused by the presence of additional stressors and intra- and interspecies variation, to give a PNEC of 0.03 mg/L or 30 µg/L. The risk quotients (PEC/PNEC) for pelagic organisms under:

1. receiving creek conditions is 0.001(0.033/30)
2. receiving lake conditions is 0.002 (0.07/30)

These risk quotients indicate a low likelihood of risk from exposures at current concentrations in the environment.

Benthic Organisms

The value selected is the maximum open lake sediment concentration in Lake Ontario of 0.0066µg/g-dw. This value was selected as the PEC. There is currently one ecotoxicological study available for benthic invertebrates exposed to PFOA, specifically the aquatic midge, Chironomus tentans (MacDonald et al. 2004). In this study, no response was observed for growth and survival at concentrations up to 100 mg/L in the exposure water. The NOEC was reported as 100 mg/L, but it should be noted that higher concentrations were not tested, and the threshold for toxicity is uncertain. The 10-day NOEC of 100 mg/L from the MacDonald et al. (2004) study was chosen as the most appropriate CTV. An application factor of 100 was applied to account for laboratory to field variations, resulting in a PNEC of 1 mg/L or 1 mg/kg. The risk quotient is 0.0066 (0.0066 mg/kg / 1 mg/kg), indicating a low likelihood of risk from exposures to benthic organisms at current concentrations.

The values derived as risk quotients for PFOA are summarized in Table 6.

Table 6. Summary of Values Derived as Risk Quotients for PFOA

In summary, the risk quotients for pelagic organisms indicate a low likelihood of risk from exposures at current concentrations in the environment. However, due to the highly persistent nature of PFOA, the potential for PFOA to affect endocrine function (including vitellogenin induction, feminization in male fish, ovary degeneration in female fish) in several species, there is the potential that concentrations in the aquatic environment may approach exposures resulting in harm in the future. The risk quotient for Canadian mammalian wildlife (i.e. polar bears) is less than 1; however, due to the persistent, bioaccumulative, and upward temporal trend of PFOA concentrations in polar bears, and the fact that other perfluoroalkyl compounds and precursors to PFOA may contribute to the overall additive or synergistic impact of PFOA in polar bears, PFOA concentrations in polar bears may approach exposures resulting in harm in the future.

Uncertainties in Evaluation of Ecological Risk

While certain data gaps and uncertainties exist, there is, nonetheless, a substantial body of information on PFOA and its precursors. For example, while the mechanism of transport of PFOA and its precursors to the Arctic is not clear, they appear to be mobile in some form, as PFOA has been measured in biota throughout the Canadian Arctic, far from known sources. Environmental pathways of PFOA to biota are not well understood, as there are relatively few monitoring data on concentrations of various precursors in air, water, effluents and sediments in Canada. While mechanisms of toxic action of PFOA are not well understood, a range of toxicological effects, including vitellogenin induction, hepatotoxicity and feminization of male fish, have been reported in a variety of species. There is limited information on the toxicology of PFOA precursors, the potential for combined or synergistic effects with PFOA, and the toxicology and potential for combined or synergistic effects of PFOA with other perfluoroalkyl acids. In addition,

analytical results from individual laboratories may not be directly comparable, according to studies by van Leeuwen et al. (2006), indicating variability in analytical results between individual laboratories.

Exposure Assessment

PFOA has been detected in a number of locations in Canada: in house dust in Ottawa, Ontario with a mean level of 106 ng/g (n=67; range 2.29 to 1234 ng/g; median 19.72 ng/g - Kubwabo et al. 2005); in tap water in Calgary, Alberta, and Vancouver, British Columbia, at 0.2 ng/L (one sample per location - Lien et al. 2006); in surface waters from US and Canadian locations on Lake Ontario and Lake Erie at levels ranging from 13 to 50 ng/L (mean of 8 samples from 4 locations on Lake Ontario: 42.5ng/L; mean of 8 samples from 4 locations on Lake Erie: 35.6 ng/L - Boulanger et al. 2004; median of 13 samples on Lake Ontario: 21 ng/L; median of 3 samples on Lake Erie: 15 ng/L – Sinclair et al. 2006); in lake water at 5 locations in the Arctic with means of 0.9 to 14 ng/L (range 0.5 to 16 ng/L – Stock et al. 2007); and in indoor air and dryer lint (2009 email from New Substances Assessment and Control Bureau, Health Canada, to Risk Assessment Bureau, Health Canada; unreferenced^[2]). PFOA contamination has also been detected in fish and water at airport sites in Canada where fire-fighting foams have been released (Moody et al. 2002; 2008 email from Transport Canada, to Bureau of Risk and Impact Assessment, Health Canada; unreferenced).

Similarly, FTOHs have been detected in a number of locations in Canada: in house dust in Ottawa, Ontario (n=59; mean 8:2FTOH: 55 ng/g – Shoeib et al. 2005), and in the air over rural and urban sites in Ontario and Manitoba (total FTOHs in Toronto, Ontario: range 113 to 213 pg/m³, mean 165 pg/m³; Winnipeg, Manitoba: range MDL-18 pg/m³, mean 11 pg/m³; Long Point, Ontario: MDL to 52 pg/m³, mean 26 pg/m³; n=3 samples per location -Stock et al. 2004; 8:2 FTOH in Toronto, Ontario: range 9-123 pg/m³, mean 55 pg/m³; n=4; Long Point, Ontario: range 25-40 pg/m³, mean 32 pg/m³; n=2 – Martin et al. 2002) and in the Arctic (total FTOHs: mean 28 pg/m³; range for 8:2 FTOH³ – Stock et al. 2007; 8:2 FTOH: range 5.8 to 26 pg/m³; n=20 – Shoeib et al. 2006).

In 3 of 54 composite food samples collected for the Canadian Total Diet Study from 1992 to 2004, PFOA was detectable, with the highest level quantified (3.6 ng/g) in a sample of microwave popcorn (Tittlemier et al. 2007). PFOA was also found in several species of fish (both raw and cooked samples; maximum level detected 1.59 ng/g) purchased at markets in Ontario in 2006 (Del Gobbo et al. 2008) and in some native foods (seal, duck and caribou; maximum level detected 0.8 ng/g) collected in Nunavut in 1997–1998 (Ostertag et al. 2009).

PFOA was detected in the packaging materials and in the vapours produced from the microwave cooking of both tested brands of prepackaged microwave popcorn, and 8:2 FTOH was found in the packaging and vapours of one of two brands of microwave popcorn. NeitherPFOA nor 8:2 FTOH was detected in the packaging materials or vapours of plain popcorn kernels cooked in a polyethylene bag (Sinclair et al. 2007). In studies that analyzed products purchased at US retail stores,PFOA was detected in new polytetrafluoroethylene-treated cookware following heating to normal cooking temperatures (Begley et al. 2005) and in water boiled in two of four new non-stick pans (Sinclair et al. 2007).PFOA and 8:2 FTOH were detected in off-gassing from all four tested brands of new non-stick cooking pans heated at normal cooking temperatures (Sinclair et al. 2007). PFOA was also detected in cooking oil following heating under normal cooking conditions in new non-stick pans in a study in Italy (Bononi and Tateo 2007). Other studies have reported no detectable PFOA following extraction of new treated cookware (Powley et al. 2005; Washburn et al. 2005; Bradley et al. 2007). Extracts from some fluorochemical-treated articles (apparel, carpet and upholstery) were found to containPFOA (Washburn et al. 2005).

PFOA was detected in 44 of 45 samples of human breast milk from Massachusetts, with concentrations ranging from < 0.0301 to 0.161 µg/mL (Tao et al. 2008), and cord blood of newborns in Canada was found to contain PFOA in three studies (Tittlemeir et al. 2004; Monroy et al. 2008; 2009 email from New Substances Assessment and Control Bureau, Health Canada, to Risk Assessment Bureau, Health Canada; unreferenced), indicating potential exposures in utero and through breast milk.

The available data suggest that Canadians are exposed toPFOA and its precursors present in the environment, including via air, drinking water and food; and from the use of consumer products, such as new non-stick cookware and PFC-treated apparel and materials such as carpets and upholstery. In general, the levels of PFOA observed in Canada in ambient air, food and drinking water are comparable to those measured elsewhere, including the United States, Europe and Asia (Fromme et al. 2009). However, in other countries, industrial sources not present in Canada have led to localized PFOA concentrations in drinking water that are much higher than the levels reported in Canada (e.g., Little Hocking, Ohio, mean concentration of 3.5 µg/L; Emmett et al. 2006a).

A quantitative estimate of exposure to PFOA based on levels in environmental media and through use of consumer products was not derived, as biomonitoring data – which aggregates exposure from all sourcse – was available. However, such estimates of total daily intake of PFOA have been published recently. Trudel et al. (2008) estimated the exposure of the general population in North America to PFOA from air, food, drinking water, dust and consumer products. The exposure estimates for low-, intermediate- and high-exposure scenarios ranged from 1 to 130 ng/kg-bw per day for all age groups. Diet was found to be the major source of exposure for the low- and intermediate-exposure scenarios, whereas consumer products (oral exposure to treated carpet, migration from treated paper into food and inhalation during treatment of clothing) were the major contributors in high-exposure scenarios. This study did not consider contributions from exposure to PFOA precursors. In another study, Fromme et al. (2009) estimated exposure to PFOA for the general population using data from North America, Europe and Japan. The upper bound of exposure was estimated to be 12.6 ng/kg-bw per day and was mainly due to dietary consumption. Exposure to FTOHs contributed less than 1% of the exposure to PFOA. This study did not look specifically at consumer product scenarios.

Biomonitoring data form the basis of this exposure assessment for PFOA, as serum levels represent aggregate exposure from all routes and sources, including exposure to precursors of PFOA. PFOA has been identified in blood samples from non-occupationally exposed populations worldwide, including in adults and newborns in Canada. The widespread detection of PFOA in human blood indicates that humans are environmentally exposed to PFOA and/or precursor compounds that degrade to PFOA. Once in the body, PFOA can bind to certain proteins (Han et al. 2003, 2005), but there is no evidence that it is modified by metabolism, conjugation or defluorination (Vanden Heuvel et al. 1991). PFOA has a relatively long half-life in humans, in the range of 2–13 years (Burris et al. 2002; Olsen et al. 2007).

PFOA has been measured in the blood of Canadian adults and newborns in several studies, at levels ranging from 0.00048 µg/mL up to 0.0072 µg/mL. Although biomonitoring data for Canada are limited, available Canadian data are consistent with data for US populations. For example, in the 2003–2004 National Health and Nutrition Examination Survey (NHANES), the geometric mean PFOA concentration in serum of 2094 Americans aged 12 years or older was 0.0039 µg/mL, with a 95th percentile of 0.0098 µg/mL. Serum PFOA levels in children and the elderly are comparable with those in adults. Table 7 shows the results from Canadian and US biomonitoring studies.

Table 7.PFOA biomonitoring Studies From Canada and the United States

Confidence in the measure of exposure (i.e., levels of PFOA in blood) is high, because although the Canadian sampling data are limited, the mean and 95th percentile values are very similar to those reported for samples collected in the United States. Additionally, the use of biomonitoring data accounts for multiple sources of exposure and eliminates the need to calculate upper bound estimates of exposure from environmental sources and consumer products and household materials.

Health Effects Assessment

Most of the relevant toxicity studies have been conducted on the ammonium salt of PFOA (APFO); studies are also available on the anion (PFO), the sodium salt (PFOA-Na) and on PFOA itself. PFOA and its salts are expected to dissociate rapidly in biological media to form the PFO anion, and the toxicological effects of exposure to PFOA, APFO and PFOA-Na are similar. Therefore, PFOA and its salts are considered to be biologically equivalent in this report. Toxicological data on individual precursors of PFOA were not reviewed as the approach for this assessment was to consider these compounds in terms of their contribution to exposure to total PFOA. A summary of the toxicological database for PFOA and its salts is provided in Appendix A. The studies considered critical to the screening assessment are those with the lowest administered doses and those that reported the lowest serum levels of PFOA associated with effects.

The lowest LOAEL identified in experimental animal toxicity studies was 0.3 mg/kg-bw per day for APFO in a short-term (14-day) oral toxicity study in rats and mice. This dose was associated with a mean serum PFOA level of 13 µg/mL and increased liver weight in male mice; and a mean serum PFOA level of 20 µg/mL and altered lipid parameters in male rats. This dose was the lowest tested; thus, no NOAEL was obtained (Loveless et al. 2006).

In a subchronic (13-week) oral toxicity study in male rats, no effects, including histopathological changes in the liver, were observed at 0.06 mg/kg-bw per day. At the next dose (0.64 mg/kg-bw per day), increased liver weight and hepatic hypertrophy were observed. These effects were not observed after an 8-week recovery period. Corresponding serum PFOA concentrations at the NOAEL and LOAEL were 7.1 and 41.2 µg/mL, respectively, at the end of the 13-week exposure period (Palazzolo 1993; Perkins et al. 2004).

Liver effects were also observed in a 14-day inhalation study in male rats. Reversible liver weight increase, reversible increases in serum enzyme activities and microscopic liver pathology, including necrosis (not reversible), occurred at exposure levels of 8 mg/m³ and above, but not at 1 mg/m³. These exposures were calculated as equivalent to doses of 2.48 and 0.31 mg/kg-bw per day, respectively^[3] (corresponding serum PFOA levels were 47 and 13 µg/mL, respectively, at the end of the 14-day exposure period) (Haskell Laboratory 1981a; Kennedy et al. 1986).

In a 26-week oral subchronic toxicity study in male monkeys, the lowest tested APFO dose of 3 mg/kg-bw per day resulted in increased liver weight (Thomford 2001b; Butenhoff et al. 2002). The average serum PFOA level based on biweekly measurements was 77 µg/mL.

In a developmental toxicity study in mice, the oral administration of APFO at 1 mg/kg-bw per day on days 1–17 of gestation resulted in increased liver weight in the dams, alterations in fetal ossification and early puberty in male pups. The mean serumPFOA level in the dams at the LOAEL of 1 mg/kg-bw per day was 21.9 µg/mL. As this was the lowest tested dose, no NOAEL was determined (Lau et al. 2006).

Two chronic toxicity studies in laboratory animals were identified. In one, CD rats were given APFO at 0, 30 or 300 ppm in the diet for 2 years (0, 1.3 or 14.2 mg/kg-bw per day for males; 1, 1.6 or 16.1 mg/kg-bw per day for females). A LOAEL of 1.3 mg/kg-bw per day was determined based on significant increases in serum liver parameters (alanine aminotransferase [ALT], alkaline phosphatase [AP] and albumin) in males. No evidence of carcinogenic activity was seen in the females, but the males showed an increase in testicular Leydig cell adenomas. This increase was significant (p = 0.05) in the high-dose group (14.2 mg/kg-bw per day) (Sibinski 1987). In the second chronic toxicity study, male CD rats were given only one dietary dose level of APFO(300 ppm; 13.6 mg/kg-bw per day) for 2 years. The exposed males had significantly higher incidences of adenomas of the liver, hyperplasia and adenomas in Leydig cells and hyperplasia and adenomas in pancreatic acinar cells (Biegel et al. 2001). Serum PFOA levels were not reported in either of the chronic studies.

Testing of PFOA or its ammonium or sodium salts (see Appendix A for specific compound details and references) for genotoxic potential produced no evidence of activity in three in vivo bone marrow micronucleus tests in mice, several Ames bacterial mutation assays and three in vitro chromosomal aberration tests (two in hamster cells and one in human cells). Positive results were obtained in one chromosome damage test in hamster cells and one micronucleus assay in human cells in vitro. PFOA caused oxidative DNA damage in human hepatoma cells in culture and in the liver of rats treated by the oral or intraperitoneal route. The genotoxicity database indicates that PFOA compounds are not mutagenic.

In rodents, PFOA triggers a peroxisome proliferation response, which is mediated by PPARa. Activation of PPAR acauses changes in liver and affects lipid metabolism and transport and other biochemical processes. The triad of benign tumours observed in PFOA-exposed male rats (liver, testicular Leydig cell and pancreatic acinar cell adenomas) is typical of PPAR agonists, including clofibrate, 2,2-dichloro-1,1,1-trifluoroethane (HCFC 123) and pirinixic acid (WY-14,643) (Cook et al. 1999; Kennedy et al. 2004). It has been proposed that the activation of hepatic PPARa, rather than direct genotoxic action, is the critical event in the induction of these tumours. The confidence in the PPARamode of action is high for liver tumours, moderate for Leydig cell tumours (LCTs) and low for pancreatic acinar cell tumours (PACTs) (reviewed in Klaunig et al. 2003).

PFOA has been shown to activate PPARa in vitro and causes hepatic peroxisome proliferationin vivo. Temporal and dose–response data support a PPARa mode of action for rat liver tumours. In the chronic dietary study by Biegel et al. (2001), PFOA caused an increase in liver weight and peroxisome proliferation in the liver of treated rats at every time point studied, starting at 1 month of treatment, and the first occurrence of liver tumours was after 12 months. In a 14-day study in male rats, Liu et al. (1996) determined a NOEL of 0.2 mg/kg-bw per day and a LOEL of 2 mg/kg-bw per day for increased liver weight and hepatic peroxisome proliferation (measured by hepatic ß-oxidation), key endpoints in liver tumour initiation. Evidence of liver carcinogenicity was observed at 13.6 mg/kg-bw per day in the chronic study by Biegel et al. (2001), but not at 1.3 mg/kg-bw per day in the chronic toxicity study by Sibinski (1987), which is in agreement with the dose–response relationship for early key events. In a subchronic study in monkeys, there was no increase in peroxisome proliferation in response to PFOA, although liver effects, including liver weight increases, were observed (Butenhoff et al. 2002).

PFOA did not induce peroxisome proliferation or cell proliferation in male rat Leydig cells (Biegel et al. 2001), suggesting a mechanism for tumour induction other than direct PPAR activation in the testes. It has been proposed that PFOA-induced LCTs in the male rat are the result of an increase in serum estradiol due to hepatic PPARaactivation and subsequent changes in activities of enzymes involved in steroid biosynthesis. Studies have shown that treatment with PFOA causes increased hepatic aromatase activity and serum estradiol levels (Liu et al. 1996; Biegel et al. 2001). In a 14-day study in male rats, Liu et al. (1996) demonstrated that hepatic aromatase activity and serum estradiol were increased only at doses that caused hepatic peroxisome proliferation. A NOEL of 0.2 mg/kg-bw per day and a LOEL of 2 mg/kg-bw per day for increased hepatic aromatase activity and serum estradiol were determined. In the chronic toxicity study by Sibinski (1987), LCTs were increased significantly at the high dose of 14.2 mg/kg-bw per day but not at the low dose of 1.3 mg/kg-bw per day. PFOA caused an increase in serum estradiol in exposed rats at 1, 3, 6, 9 and 12 months, prior to the first occurrence of LCTs, in the chronic dietary study by Biegel et al. (2001). Thus, the proposed PPARa mode of action is consistent with dose–response and temporal associations with increased serum estradiol. Within the testis, estradiol modulates growth factor expression, which could lead to hyperplasia and adenomas (Biegel et al. 1995). PFOA and other PPARa agonists, including bezafibrate, monoethylhexyl phthalate (MEHP) and WY-14,643, have also been shown to inhibit testosterone production in Leydig cells in vitro (Klaunig et al. 2003), although serum testosterone was not decreased in the 2-year rat bioassay with PFOA (Biegel et al. 2001). In non-human primates, treatment with PFOA did not cause an increase in serum estrogen or a decrease in serum testosterone, and no abnormal histopathology was noted in the testes in the 26-week study (Butenhoff et al. 2002).

A hepatic PPAR-mediated pathway may also be involved in the induction of PACTs in male rats, although the confidence in this mode of action is lower than for liver tumours or LCTs (reviewed in Klaunig et al. 2003). In vivo mechanistic evidence is mainly indirect, from other PPAR-agonists that have been demonstrated to cause rat PACTs. Pancreatic acinar cell (PAC) hypertrophy, hyperplasia and adenomas in rat have been shown to be modified by steroid hormones, including estradiol and cholecystokinin. PFOA exposure could cause reduced bile flow or changes in bile composition resulting from downstream effects of hepatic PPARa activation. This may, in turn, cause an increase in cholecystokinin (CCK) levels and stimulation of the PACs . Although no mechanistic data are available for PFOA itself, decreased bile flow and bile acid concentration as well as a sustained increase in plasma CCK occurred in rats treated with the PPARaagonist WY-14,643. These changes were observed as early as 3 months, earlier than the time point at which PACTs were first observed. PFOA-treated rats had increased pancreatic cell proliferation at 15, 18 and 21 months, and acinar hyperplasia was significantly increased in the chronic dietary study by Biegel et al (2001). No abnormal histopathology was noted in the pancreas of PFOA-treated monkeys in the 26-week study (Butenhoff et al. 2002).

In a draft risk assessment, the US EPA stated that “there is strong evidence to conclude that the liver toxicity and liver adenomas that are observed in rats following exposure toPFOA result from a PPARa-agonist mode of action,” which is unlikely to occur in humans. They also conclude that although the LCTs and PACTs may be relevant to humans, they probably do not represent a significant cancer hazard due to quantitative differences in receptor expression and other toxicodynamic factors (US EPA 2005). However, the Science Advisory Board reviewed the US EPA’s risk assessment and concluded that there may be other modes of action for liver tumours and that as the modes of action for LCTs and PACTs are unknown, they should be considered relevant to humans (US EPA 2006b).

Recent studies in mice lacking expression of the PPARa gene (PPARa-null) suggest that some of the effects associated with exposure to PFOA are independent of the peroxisome proliferation pathway. Developmental effects, such as delayed eye opening, deficits in postnatal weight gain and reduced postnatal survival, were observed in wild-type but not PPARa-null mice, suggesting that these effects are dependent on PPARa expression. However, early pregnancy losses were observed in both strains of mice, indicating involvement of pathways other than peroxisome proliferation (Abbott et al. 2007). Yang et al. (2002) noted that reductions in spleen weight and number of splenocytes in wild-type mice treated with PFOA in the diet for 7 days did not occur in treated PPARa-null mice. PFOA-exposed mice also had reduced thymus weight and number of thymocytes, effects that were attenuated in the PPARa-null mice. However, significant liver weight increase occurred in both wild-type and PPARa-null mice. Gene profiling studies also suggest that PFOA can alter mouse liver genes independent of PPARa (Rosen et al. 2008a, b);however, the toxicological relevance of this is not known.

Available epidemiological studies on the adverse health effects of exposure to PFOA include cross-sectional general population studies, studies in populations exposed to higher levels of PFOA through contaminated drinking water and occupational exposure studies. However, no causal relationships between PFOA and adverse health effects have been established due to numerous confounding factors including exposure to multiple chemicals.

Recently, two studies (one cross-sectional study in the United States and one cohort study in Denmark) have given limited suggestions of a weak association between gestational exposure to PFOA and reduced birth weight (Apelberg et al. 2007b; Fei et al. 2007). However, the magnitude of the effect was small, given normal variations in the parameters measured, and all of the children were within the normal range of variation. No associations between maternal serum PFOA levels and birth weight were observed in other general population studies in Canada (Monroy et al. 2008; Hamm et al. 2009), Japan (Washino et al. 2009) or a US community with a mean serum PFOA level 10 times higher than that of the general US population (Stein et al. 2009).

Confidence in the effects assessment for PFOA is moderate to high, as the toxicological database covers a wide range of endpoints and life stages, several species and both sexes.

Characterization of Risk to Human Health

The available data are sufficient to derive margins of exposure (MOEs) based on comparisons of serum PFOA levels in laboratory animals at the critical effect levels with serum PFOA levels in humans from biomonitoring studies. The MOEs between the serum PFOA levels associated with the most sensitive effects in laboratory animals and the serum PFOA levels in adult Canadians range from approximately 3800 to 22 600 for mean serum values and from 2100 to 12 600 for conservative 95th percentile values. Using 95th percentile serum concentrations for US adults or children, MOEs range from about 1300 to 7900 (Table 8).

Table 8. Margins of Exposure

The experimental animal studies that were used for generation of MOEs were those with the lowest administered doses and serum PFOA levels associated with effects. A 14-day oral study in mice and rats (Loveless et al. 2006) had the lowest LOAEL identified in any study (0.3 mg/kg-bw per day), and MOEs were calculated based on the serum PFOA levels in mice and rats associated with this dose. Although no NOAEL was determined in the critical study, there is supporting evidence from other studies, such as a 13-week oral rat study and a 14-day rat inhalation study, in which no effects were observed at 0.06 and 0.31 mg/kg-bw per day, respectively (Haskell Laboratory 1981a; Kennedy et al. 1986; Palazzolo 1993; Perkins et al. 2004).

MOEs were also calculated based on serum levels in mice dosed at the LOAEL for developmental toxicity (1 mg/kg-bw per day) (Lau et al. 2006). Although no NOAEL was determined in this study, in a follow-up study designed to examine the mechanism of action of PFOA-induced developmental effects, a NOAEL for developmental toxicity of 0.3 mg/kg-bw per day was determined, based on decreased neonatal survival at 0.6 mg/kg-bw per day and delayed eye opening at 1 mg/kg-bw per day. As serum PFOA levels were not measured until 22 days post-partum, effect levels from this study were not used for determining an MOE (Abbott et al. 2007).

A 26-week toxicity study in monkeys was also selected for MOE estimation, as primates are considered to be a better surrogate for humans than are rodents (Thomford 2001b; Butenhoff et al. 2002). The steady-state serum PFOA level of monkeys dosed at the LOAEL of 3 mg/kg-bw per day was used for the calculations. No NOAEL was determined in this study.

The use of serum levels for the MOE calculations significantly reduces uncertainties associated with interspecies and intraspecies differences in pharmacokinetics. In laboratory animals, PFOA distributes primarily to the serum and liver (Vanden Heuvel et al. 1991; Butenhoff et al. 2004a; Hundley et al. 2006; Kudo et al. 2007). Data on PFOA tissue distribution in humans are limited, but it is assumed that the pattern would be similar to that observed in experimental animals (including male rats, male and female mice and non-human primates) (Hundley et al. 2006). PFOA has been detected in human post-mortem liver samples in two studies (Olsen et al. 2003b; Maestri et al. 2006). Since data on PFOA levels in human tissues are limited, the availability of data on PFOA concentrations in serum makes it the most appropriate measure of internal exposure. Due to the long half-life of PFOA and its lack of metabolism, human adult serum levels of PFOA represent cumulative (lifetime) exposure to PFOA and any contributions from its precursors. Serum levels of PFOA are comparable in children, adults and the elderly. Although the critical toxicology studies in experimental animals were conducted over less than lifetime exposure durations, pharmacokinetic data indicate that the reported serum PFOA levels would represent steady state (Vanden Heuvel et al. 1991; Butenhoff et al. 2004a; Lau et al. 2006).

Where an MOE is estimated on the basis of effects in rodents, the possible role of peroxisome proliferation should be acknowledged. Rats and mice are sensitive to the effects of peroxisome proliferators, whereas monkeys and humans are relatively non-responsive at similar doses (reviewed in Klaunig et al. 2003; Kennedy et al. 2004). As humans are generally less susceptible than rodents to peroxisome proliferators, MOEs based on peroxisome proliferation–dependent effects in rodents would be conservative. As the margins were calculated based on the most sensitive effects and species, they are considered protective of both PPAR-dependent and PPAR-independent effects.

APFO has induced tumours in exposed rats, but PFOA compounds have not been tested for carcinogenic potential in any other laboratory animal species. The PPAR-agonist mode of action proposed for rat liver, testes and pancreatic tumours may not be relevant for humans, but human relevance has not been definitively determined according to established frameworks (Meek et al. 2003; Boobis et al. 2006).

PFOA has been shown to activate the human PPARa in cell culture (Takacs and Abbott 2007; Wolf et al. 2008).PPARaactivation has a wide range of effects, including regulating the expression of genes involved in cell growth and survival. In fact, some PPARa ligands have been shown to possess anti-tumorigenic properties, such as suppression of growth of several types of human cancer cells in vitro and inhibition of carcinogenesis in vivo (reviewed in Pozzi and Capdevila 2008), makingPPARa a potential target for cancer therapy. It should also be noted that other PPARa ligands, such as fibrates, which cause a high incidence of tumours in rodents, are commonly used therapeutically in humans, with no evidence of carcinogenicity in epidemiological studies (reviewed in Peters et al. 2005).

Although the modes of action for tumour induction have not been demonstrated conclusively and the relevance of the rat tumours to carcinogenicity in humans is uncertain, the genotoxicity database indicates that PFOA is not directly mutagenic. Thus, as the tumours observed in male rats are not considered to have resulted from direct interaction with genetic material, a threshold approach is used to characterize risk to human health. The MOEs generated for non-neoplastic effects are protective with respect to the increased incidence of benign tumours observed in the chronic studies of PFOA in rats, since a) the tumours were observed only at PFOA doses higher than those that induced non-neoplastic effects, b) the genotoxicity database indicates that PFOA is not mutagenic, c) rat liver tumours are likely induced via a mode of action that is not relevant to humans and d) there were no PFOA-related effects in non-human primates that have been associated with the development of pancreatic or testicular tumours in rats.

Uncertainties in Evaluation of Risk to Human Health

The use of serum levels reduces the uncertainty associated with a determination of the upper-bound estimate of human intake of PFOA, owing to the limited available data on levels of PFOA and precursors in air, foodstuffs, drinking water and breast milk and resulting from contact with household materials treated with perfluorinated substances. Moreover, the levels of PFOA in human serum provide a measure of aggregate exposure from multiple sources and exposure routes. The use of serum levels also significantly reduces uncertainties associated with interspecies and intraspecies differences in pharmacokinetics.

Use of 95th percentile serum values is conservative. Furthermore, measures have recently been taken to reduce global facility emissions and product content of PFOA and related chemicals, including in Canada (US EPA 2006c; EPA 2009). Two recent biomonitoring studies in the United States have already shown declining serum PFOA levels (Calafat et al. 2007; Olsen et al. 2008). Exposure to PFOA in Canada is also expected to decrease over time, although serum data are currently not available to confirm this. The MOEs presented here are based on serum levels from 1994 to 2004, and the MOEs may increase over time as exposure levels are predicted to decrease in the future.

Uncertainty remains regarding the mode of action for tumour induction; however, as the database for genotoxicity suggests thatPFOA is not mutagenic, the MOEs based on non-neoplastic effects in the most sensitive species are considered protective of any potential carcinogenic effects in humans.

Footnotes

Based on the persistent nature of PFOA in all media in the environment, its potential to bioaccumulate and biomagnify in terrestrial and marine mammals, the fact that environmental concentrations may approach effect levels of concern, the temporal trends of PFOA in polar bears, the widespread occurrence of PFOA in biota, including in remote areas, and the fact that other perfluoroalkyl compounds and precursors to PFOA may contribute to the overall additive or synergistic impact of PFOA in biota, it is proposed that PFOA, its salts and its precursors are entering the environment in a quantity or concentration or under conditions that have or may have an immediate or long-term harmful effect on the environment or its biological diversity. While there is scientific evidence that PFOA and its salts can accumulate and biomagnify in terrestrial and marine mammals, PFOA and its salts do not meet the criteria for bioaccumulation as defined in the Persistence and Bioaccumulation Regulations. Additionally, it is proposed that PFOA and its salts meet the criteria for persistence as set out in the Persistence and Bioaccumulation Regulations.

Based on a comparison of upper-bound levels of PFOA in the blood (serum) of humans with serum levels that are associated with the development of adverse effects in laboratory animals, it is considered that the resulting margins of exposure are adequately protective of human health. It is proposed that PFOA and its salts are not entering the environment in a quantity or concentration or under conditions that constitute or may constitute a danger in Canada to human life or health. Precursors of PFOA were not individually assessed, but the approach for the assessment included the contribution of precursors to total PFOA, the moiety of toxicological concern, as they can degrade to PFOA in the environment.

It is therefore proposed that PFOA, its salts and its precursors meet one or more of the criteria set out in section 64 of CEPA 1999.

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Table. Additional Studies