Screening Assessment - Canada.ca (2024)

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Environment Canada
Health Canada

August 2012

(PDF Version - 599 KB)

• Synopsis
• Introduction
• Substance Identity
• Physical and Chemical Properties
• Sources
• Uses
• Releases to the Environment
• Environmental Fate
• Persistence and Bioaccumulation Potential
• Potential to Cause Ecological Harm
• Potential to Cause Harm to Human Health
• Conclusion
• References
• Appendix 1: Summary of health effects information for PFOA

Synopsis

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^[1] 335-67-1, its salts and its precursors under sections 68 and 74 of CEPA 1999. The ammonium salt of PFOA and some precursors to PFOA, present on Canada’s Domestic Substances List (DSL), were categorized under section 73 of CEPA 1999. While PFOA itself is not on the DSL, PFOA can be formed in the environment through transformation or degradation from a variety of other perfluorinated chemicals. PFOA was identified for assessment based on 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). In this assessment, 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. PFOA and its salts are used as polymerization aids in the production of fluoropolymers and fluoroelastomers. PFOA, itself, is not manufactured in Canada; however, quantities of the ammonium salt are imported.

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. 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. 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 found in commercial products and detectable in the atmosphere and can degrade or transform to PFOA through biotic or abiotic pathways.

Once in the environment, PFOA is extremely persistent and not known to undergo significant further abiotic or biotic degradation under relevant environmental conditions. PFOA is highly soluble in water 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 is likely attributable to the long-range transport of PFOA (e.g., via ocean currents) and/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 freshwater (ND–11.3 µg/L) and freshwater sediments (0.3–7.5 µg/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 turbot 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 an accidental release of fire-fighting foam in Etobico*ke Creek (Ontario), PFOA was measured in common shiner liver at a maximum concentration 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 (1972 – 2002 and 1984 – 2006) and sea otters (1992 – 2002). 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, PFOA is present mainly in its ionic form in environmental media. Due to the perfluorination, the perfluorinated chains are both oleophobic and hydrophobic. PFOA primarily binds to albumin proteins in the blood of biota and, as a result, is present in blood and highly perfused tissues such as liver and kidney, rather than lipid tissue. The numeric criteria for bioaccumulation, outlined in the Persistence and Bioaccumulation Regulations of CEPA 1999, are based on bioaccumulation data for freshwater aquatic species (fish) only and for substances that preferentially partition to lipids. As a result, the criteria may not completely reflect the bioaccumulation potential 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 non-aquatic 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 of embryos. However, PFOA accumulated in the liver of these embryos to concentrations 2.9 – 4.5 times 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 20 mg/L increased the chemosensitivity in marine mussels. PFOA at 25.9 mg/L activated the mammalian peroxisome proliferator–activated receptor α (PPARα) in the livers of Baikal seals--PPARα 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. In 2-year carcinogenicity bioassays in rats, males administered a high dose of PFOA ammonium salt (APFO) in the diet had significantly higher incidences of adenomas of the liver hepatocytes, Leydig cells in the testes and pancreatic acinar cells. 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. There is some evidence to suggest that PFOA may be capable of causing indirect oxidative DNA damage.

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 or plasma) 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 or plasma levels found in non-occupationally exposed adults, infants, and children (0.00162–0.0195 µg/mL) results in margins of exposure greater than 660. These margins are considered to be adequately protective to account for uncertainties in the hazard and exposure databases.

The ecological assessment is based on a weight-of-evidence approach regarding persistence, bioaccumulation, temporal trends in some species (i.e., the polar bear), long-range transport and the widespread occurrence and concentrations of PFOA in the environment and biota (including remote areas of Canada). Based on the available information, it is concluded that PFOA, its salts and its precursors are entering or may be 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. In addition, it is concluded that PFOA and its salts meet the criteria for persistence as set out in the Persistence and Bioaccumulation Regulations. PFOA and its salts do not meet the criteria for bioaccumulation as set out in the Persistence and Bioaccumulation Regulations. Nevertheless, the weight of evidence is sufficient to conclude that PFOA and its salts accumulate and biomagnify in terrestrial and marine mammals.

Based on the available information on the potential to cause harm to human health and the resulting margins of exposure, it is concluded 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.

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

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.

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Introduction

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. 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.

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 are on the Domestic Substances List (DSL) and were found to meet the ecological categorization criteria for persistence and/or bioaccumulation potential and/or 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 February 2011 for ecological sections of the document and up to April 2011 for the human health sections of the document. 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 relevant scientific and technical information and develop conclusions based on multiple lines of evidence, including persistence, exposure, trends, 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^[2].

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. Additionally, the draft of this screening assessment was subject to a 60-day public comment period. While external comments were taken into consideration, the final content and outcome of the screening 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.

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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 (OECD TG 301C) 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 also 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 in the Canada GazetteNotice with respect to certain perfluoroalkyl and fluoroalkyl substances, on the Canada GAzette, Part I: Vol. 139, No. 3. 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.

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Physical and Chemical Properties

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 pKa value commonly reported and used is approximately 2.5 (Kissa 1994). However, Burns et al. (2008) determined a pKa of 3.8 ± 0.1, suggesting that the neutral species can exist in the environment. Jasinski et al. (2009) determined a pKa of 3.3 ± 0.4. Goss and Arp (2009), using analogues and molecular models, suggested that the pKa ranges from 0 to 4. Goss (2008) suggested that PFOA is expected to have a low pKa, 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.

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

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Sources

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, hom*ologues and by-products (US EPA 2002). The perfluorooctanoyl fluoride is then separated and hydrolyzed to yield to yield a mix of ~70% linear and various branched isomers of PFOA (Riddell et al. 2009). 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% linear 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.

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Uses

Releases to the Environment

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 no removal of PFOA during the sewage treatment plant processes. Precursors such as FTCAs and FTOH degradation products have been measured in influent and primary treatment samples, but not in secondary treatment waters (Sinclair and Kannan 2006).

Boulanger et al. (2005) conducted a lakewide mass budget analysis of eight perfluorooctane surfactants, including PFOA, 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 telomer alcohols. The authors suggested that the residual fluorinated telomere alcohol contribution to the atmospheric load of fluorinated telomer alcohols is significant and may be the dominant source, given that the release of these residual fluorinated telomer 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 (i.e., garbage dumps, open pit burning dumps and metal waste dump) (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. 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). Based on the average concentration of PFOA, the annual estimated total loading in leachate is 0.1 kg (Conestoga-Rovers & Associates 2011). 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. The 6:2 FTOH, 8:2 FTOH, and 10:2 FTOH ( i.e., precursors to PFOA) have been measured in the air at two landfill sites in Ontario; concentrations were < 5000 pg/m³ (upwind) and < 25 000 pg/m³ for on-site (Ahrens et al. 2010b).

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 8:2 PAPS via cleavage of the phosphate ester linkage, releasing free 8:2 FTOH, with subsequent biotransformation to PFOA. Lee et al. (2010) showed that microbial mediated biodegradation products (i.e. 8:2 FTOH and 10:2 FTOH) from 8:2 mono substituted PAPS and 10:2 monosubstituted PAPs (which may be present in commercial products) may be present in WWTP influent and effluent.

Wallington et al. (2006) used a three-dimensional global atmospheric chemistry model (IMPACT) to indicate that n-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 and rainbow trout (Hagen et al. 1981, Butt et al. 2010b). A β-oxidation mechanism was proposed by Hagen et al. (1981) to account for the formation of PFOA in rats. Butt et al. (2010a) found that exposure to 7:3 fluorotelomer saturated carboxylate (FTCA) did not result from the formation and accumulation of PFOA in rainbow trout; however, PFOA was formed in the 8:2 FTCA and 8:2 fluorotelomer unsaturated carboxylate (FTUCA). The author proposed a beta-oxidation pathway proceeding from 8:2 FTUCA ? 7:3 β-keto acid ? 7:2 ketone ? PFOA or that PFOA could be formed directly through the β-oxidation of the 7:3 keto acid. Butt et al. (2010b) exposed rainbow trout to 8:2 fluorotelomer acrylate (FTAc) via dietary exposure and found that 8:2 FTCA, 8:2 FTUCA and 7:3 FTCA were formed; however, the overall formation and accumulation of PFOA was low. Frömel and Knepper (2010) suggested that ultimate ethoxylate shortening of fluorotelomer ethoxylates (FTEO) may result in FTOHs and thus contribute as a potential source of perfluorinated compounds such as PFOA.

Dinglasan et al. (2004) showed aerobic biodegradation 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 β-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 of ¹⁴C-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 β-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 β-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. 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).

Yoo et al. (2010) measured FTOHs in soil from fields (near Decatur, Alabama) to which sewage sludge had been applied. Sludges generated at a WWTP in Decatur, Alabama have been applied to agricultural fields for more than decade - this WWTP received waste streams from industries that work with fluorotelomer compounds (Washington et al. 2010). Yoo et al. (2010) found that sludge-amended fields had surface soil FTOH concentrations ranging from 5 to 73 ng/g dry weight. The highest FTOH concentration was for 10:2 FTOH which had concentrations ranging from <5.6 to 166 ng/g dry weight. The half-lives for FTOHs ranged from 0.85 to 1.8 years suggesting that sludge application is a possible pathway for the degradation of FTOHs to PFOA and other perfluorocarboxylic acids. Washington et al. (2010) also found that these sludge-amended fields have high concentrations of perfluorocarboxylic acids, including PFOA (<320 ng/g dry weight).

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 hom*ologous 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 produce PFOA.

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 branched PFOA 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).

Gewurtz et al. (2009) found PFOA as well as 8:2 FTUCA and 10:2 FTUCA in window film concentrations from indoor/outdoor/downtown/suburban/rural/carpet stores locations in Toronto, Ontario. Nilsson et al. (2010) found perfluorinated compounds, including PFOA in the blood of ski wax technicians (PFOA concentrations ranged from 4.8 to 535 ng/L). Nilsson et al. (2010) suggested that fluorinated organic compounds are added to glide waxes to prevent adhesion of snow, ice and dirt; fluorinated ski waxes are applied using heat of approximately 130-220 °C where airborne particles and fumes containing a blend of gaseous organochlorine compounds are emitted. However, the authors did not analyze the glide waxes, themselves, to determine the presence of perfluorinated compounds.

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Environmental Fate

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. Solubility is dependent on the acid dissociation constant (pK_a) of the acid form. Goss (2008) suggested that PFOA is expected to have a low pKa, such that > 99% of the compound will occur in its anionic form (i.e., PFO) under most environmental conditions. As such, modelling results presented in peer-reviewed literature for global fate, transport/partitioning and bioaccumulation should be viewed with caution, as they may not have accounted for the pH of the environmental compartment in conjunction with the pKa values of the compound. The ammonium salt of 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 evident by its detection in this medium (Giesy and Newsted 2001; Stock et al. 2007). A comparison of concentrations observed in the aqueous phase and in sediments suggests that sediments are unlikely to be a major sink for PFOA (Oakes et al. 2004; Masunaga and Odaka 2005). However, 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 L/kg. Organic carbon adsorption coefficient (K_oc) values ranged from 48.8 to 229 L/kg, and organic matter adsorption coefficient (K_om) values ranged from 28.4 to 133 L/kg. These values indicate that PFOA is more likely to sorb to organic carbon in soils than to other soil solids. Moody and 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).

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Persistence and Bioaccumulation Potential

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 with respect to OH radicals 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. The presence 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; Yamash*ta et al. 2005) or volatile precursors to PFOA through the atmosphere (Stock et al. 2007). Webster and Ellis (2010) proposed sea spray as a mechanism for the generation of PFOA in the gas phase from PFO in a water body which has the potential to contribute large amounts of PFOA to the atmosphere and so significantly contribute to the concentrations measured in remote areas.

There is a possibility that the presence of perfluorinated compounds in the Canadian Arctic may be due, in part, to the presence of former military bases (Iqaluit, Sarcpa Lake, Resolution Island) and/or Distant Early Warning Line (DEW) stations (northern shores of Alaska to Cape Dyer on the Baffin Island coast) that may have used perfluorinated-based products (Stow et al. 2005; Poland et al. 2001). These sites are considered remediated, but perfluorinated compounds such as PFOA were not measured or analyzed, with available literature (Stow et al. 2005; Poland et al. 2001) only identifying the concentrations of polychlorinated biphenyls (PCBs) and metals during the remediation process. PFOA has been detected in Arctic landfills at 22 – 1083 ng/g (i.e. garbage dumps, open pit burning dumps and metal waste dump) (Ikonomou 2006) suggesting sources from local consumer/industrial activities. Nonetheless, PFOA has been measured in biota and various environmental media in remote Arctic areas far from local or regional sources.

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 using pKa between 2 – 3 at pH 4 and 7(Prevedouros et al. 2006). Armitage et al. (2006) estimated a net PFO flux of between 8 and 23 tonnes per year to the Arctic using pKa between 2.5 – 2.8 at pH 5-8. 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. 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).

In summary, PFOA has the potential for long-range transport and has been shown to be persistent, with studies indicating no abiotic or biotic degradation in the environment under relevant environmental conditions. Therefore, based on the empirical and physical-chemical properties, PFOA and its salts meet the persistence criteria in water, soil, sediment and air (half-lives in soil and water ≥ 182 days and half-life in sediment ≥ 365 days; half-life in air ≥ 2 days or evidence of atmospheric transport to remote regions such as the Arctic) as set out in the Persistence and Bioaccumulation Regulations (Canada 2000).

PFCAs, including PFOA, have the combined properties of oleophobicity, hydrophobicity, and hydrophilicity over different portions of these molecules. The carboxylate functional group attached to the perfluorinated chain, for example, imparts polarity to the molecule. Due to these properties, the assumption that the hydrophobic and lipophilic interactions between compound and substrate are the main mechanisms governing partitioning are considered not applicable for PFOA. Although the K_ow has been calculated or modelled for the neutral form of PFOA and its various salts (see Table 2), current literature suggests that the K_ow is a problematic parameter for ionized surfactants because of their tendency to aggregate at the interface of a liquid–liquid system. Consequently, the K_ow has not been considered a reliable indicator of bioaccumulation potential for perfluorinated substances. However, Webster and Ellis (2011) are stating that PFCAs, including PFOA, are not surface-active, and that the K_ow is a predictor of both lipid and protein partitioning in biota. Therefore, the authors are stating that the equilibrium distribution models for the hydrophobic neutral partitioning of perfluorocarboxylic acids, including PFOA, are applicable.

Regulatory criteria (bioconcentration factors [BCFs] and bioaccumulation factors [BAFs]) have been developed under CEPA 1999 (Canada 1999) to determine whether a substance is to be considered bioaccumulative. However, these threshold criteria are based on historical experience with neutral, non-metabolized organic substances. These criteria, based on the Federal Toxic Substances Management Policy (TSMP) persistence and bioaccumulation criteria,were developed in the mid-1990s and formally published in 1995 (Canada 1995b). The criteria identify lipophilic substances with the potential to bioaccumulate, primarily in freshwater aquatic systems (fish). Substances that meet the criteria, i.e., BAF or BCF ≥ 5000 or log K_ow ≥ 5, have significant potential for bioaccumulation at the organism level and biomagnification through the food web. However, information on BAFs, BCFs or log K_ows is but one part of the overall weight of evidence in determining the overall potential of a substance to accumulate in organisms. Furthermore, a substance may be deemed to be sufficiently bioaccumulative to cause concern, even if regulatory criteria are not met.

Given the number of available experimental studies, the emphasis in this assessment has been placed on the results of experimental bioaccumulation and biomagnification studies.

PFOA is absorbed in juvenile rainbow trout (Oncorhynchus mykiss), with a BCF of 4.0 (Martin et al. 2003b). Dietary exposure to PFOA in the same species did not result in PFOA bioaccumulation (BAF = 0.038) (Martin et al. 2003b). 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 non-aquatic 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 is 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 (Kudo and Kawashima 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 serum 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 Canadian Arctic 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 δ¹⁵ 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:

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 δ¹⁵ 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 only 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 hom*ogenates 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). Martine van den Heuvel-Greve et al. (2009) found BMFs of 3.8 and 23, respectively, for the benthic and the pelagic food webs with harbour seals (Phoca vitulina) as the apex predator. BMFs for other species in the Westerschelde, Netherlands estuary ranged from 0.03 to 31.

Jeon et al. (2010a) studied the effects of salinity on the whole-body bioaccumulation of perfluorinated compounds, including PFOA, in the Pacific oyster (Crassostrea gigas). With increasing salinity (10 -34 practical salinity units (psu)), the BCFs fluctuated from 0.8 to 3.0 whereas the BAFs increased from 9.6 to 19.4. The authors suggest that the increased accumulation is mainly due to the increase in dietary uptake due to a possible altered physiology of oysters with changing salinity. The authors also suggest that there is an enhanced sorption of perfluorinated compounds to particulate matter which can increase the risks to benthic organisms and filter-feeding bivalves.

Jeon et al. (2010b) determined the serum and liver BCFs of perfluorinated compounds, including PFOA, on blackrock fish (Sebastes schlegeli) at varying salinities (10, 17.5, 25, and 34 psu). The bioconcentration at 34 psu was greater for the perfluorinated compounds except for PFOA than at other salinities. The serum BCFs ranged between 578 (at 10 psu) to 357 (at 34 psu). The liver BCFs ranged from 73 to 93.

Kwadijk et al. (2010) determined the BAFs between water and eel (30-45 cm fillets) (Anguilla anguilla) from 21 locations in The Netherlands. The water-eel BAF for PFOA was calculated to be 1.09 and the BAF for the linear isomer of PFOA was 1.12.

Table 5 provides a summary of the available bioaccumulation data.

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. When the potential for toxicity in consumer organisms is being determined, it is also 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). As perfluorinated substances partition to liver and blood, the site of toxic action is often considered to be the liver. Thus, from a toxicological perspective, bioaccumulation (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.

Bioaccumulation may also indicate either direct exposure in organisms that have accumulated PFOA, or indirect exposure in organisms that consume prey containing PFOA (via food chain transfer). This is especially true for organisms at the higher trophic levels (e.g., polar bear), where whole-body analysis is not feasible. While it is feasible to measure whole-body bioaccumulation in smaller species at lower trophic levels, the lower trophic status of the organism means that the estimated overall bioaccumulation for perfluorinated substances may be underestimated. BCFs and (particularly) BMFs based on concentrations in whole organisms may provide a useful measure of overall potential for transfer up the food chain. 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.

As such, it is considered that based on whole-body data, PFOA has low to moderate potential to accumulate in aquatic species. However, based on organ-specific data, PFOA may be considered to accumulate and biomagnify in terrestrial and marine mammals. Therefore, based on the available empirical values, PFOA and its salts do not meet the bioaccumulation criterion (BAF or BCF ≥ 5000) as set out in the Persistence and Bioaccumulation Regulations (Canada 2000a). However, there is scientific evidence that PFOA and its salts accumulate and biomagnify in terrestrial and marine mammals.

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Potential to Cause Ecological Harm

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 in lake water

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 in particulates 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

Moody et al. (2002) measured PFOA concentrations in surface waters in Etobico*ke Creek, a tributary to Lake Ontario, after an accidental release of aqueous fire-fighting foam. The malfunction released 22 000 L of fire retardant foam and 450 000 L of water from the sprinkler system into storm sewers leading to Spring Creek (approximate travel distance of 1.8 km) and subsequently to Etobico*ke Creek, which empties into Lake Ontario. Background concentrations were also measured at an upstream site. Results indicated downstream PFOA concentrations up to 11.3 µg/L. PFOA was also detectable at the upstream site at lower concentrations (ND–0.033 µg/L).

Scott et al. (2009) sampled 38 rivers (upstream and downstream of populated areas) across Canada for perfluorinated compounds during 2001 to 2008. Concentrations of PFOA ranged from 4.4 x 10^-5 to 0.0099 µg/L. The maximum concentrations of PFOA occurred from Lake Erie’s tributaries through to the St. Lawrence River. Most sites downstream of urban areas had higher concentrations than upstream sites. Background sites, such as glacial meltwater from British Columbia, had lower concentrations of PFOA.

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 0.011µg/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. Ahrens et al. (2010b) collected surface water samples from November 2 to December 30, 2008 along the latitudinal gradient from the North Sea (northern Europe) and Antarctica. PFOA was detected in 79% of the dissolved phase samples ranging in concentration from <5.2 x 10^-6 to 0.000223 µg/L. PFOA was not detected in the particulate phase. 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) (Yamash*ta et al. 2005). PFOA was detected at concentrations ranging from 0.000 001 5 to 0.000 192 µg/L, followed by PFOS at 1.1 x 10^-6–0.0577 µg/L. Yamash*ta 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.

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.

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 Etobico*ke 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 hom*ogenates 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) (0.000493–0.00814 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 and 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., at 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 0.00006 to 0.00657 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 0.00033 to 0.00820 µg/mLfor the red panda and from 0.00032 to 0.00156 µg/mLfor 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 0.00004–0.00018 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 0.00005 to 0.0018 µg/mL (Gulkowska et al. 2005). The study also measured whole-blood PFOA concentrations in cod, which ranged from 0.00005 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. 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.

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). Temporal trends in PFOA concentrations were not found in archived peregrine falcon (Falco peregrinus) eggs from 1974 to 2007 as PFOA was not detected. It should be noted that due to the low breeding success in the first 20 years, only a few eggs were collected and no eggs were collected between 1987 to 1991. Therefore, all eggs up to 1999 were analyzed individually. Eggs were pooled from the year 2000 and onward. (Holmström et al 2010).

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). O’Connell et al. (2010) analyzed 163 juvenile loggerhead turtle (Caretta caretta) plasma and serum for spatial and temporal trends of perfluorinated compounds. The turtles were captured within 8.4 km of the nearest shore from Charleston (South Carolina), Cape Canaveral (Florida), Core Sound (North Carolina), Chesapeake Bay (Maryland) and Florida Bay (Florida). PFOA plasma/serum concentrations ranged from < 7.6 x 10^-5 to 0.000993 µg/g. Spatially, Florida Bay turtles accumulated more PFOA compared to other PFCAs suggesting local sources for turtles residing in Florida Bay. Temporal trends were examined over a nine year period in loggerheads captured near Charleston and concentrations of PFOA did not change significantly through time.

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).

The most sensitive pelagic organism was found to be the freshwater alga, Pseudokirchneriella subcapitata. The 96-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; Ward et 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).

Li (2008) conducted seed germination and 5-day 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 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.

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 17β-estradiol for 48 hours significantly inhibited 17β-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 17β-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.

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 (20 mg/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 α (PPARα) in the livers of Baikal seals--the first reported identification of PPARα complementary deoxyribonucleic acid (DNA) in an aquatic wildlife species. PPAR is a member of the ligand-activated nuclear hormone receptor superfamily. PPARα 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). Yang (2010) exposed male Japanese medaka (Oryzias latipes) 10, 50, and 100 mg/L PFOA for 7 days. There were no impacts on survival, relative live and gonad size, and condition factor at any concentration. However, PFOA induced inhibition of catalase activity at high doses with no changes of superoxide dismutase or glutathione peroxidase activities in the liver suggesting that PFOA may induce peroxisomal fatty acid oxidation and impose oxidative stress through the alteration of cellular oxidative homeostasis in the liver. PFOA also increased the mRNA levels of pro-inflammatory cytokines suggesting that it may be involved in inflammation and tissue injury. The effects on peroxisome proliferation and cytokine expression suggest that chronic exposure to PFOA can be carcinogenic in the Japanese medaka – peroxisome proliferation is a key carcinogenic mechanism and chronic inflammation is associated with an increased risk of cancer in rodents.

The potential impact of exposure to perfluorinated compounds on liver lesions was investigated in East Greenland polar bears (Sonne et al. 2008). Liver parameters examined 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). Liver samples were also analyzed for several perfluorinated compound including C9, C10, C11, C12 and C13 PFCAs. 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. However, in another study by Flanary et al. (2010), no correlations were found between any perfluorinated compound, including PFOA, and blood chemistry parameters (e.g. cholesterol, creatinine, albumin, total serum ion etc.) for the northern fur seal (Callorhinus ursinus).

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).

The approach taken in this ecological screening assessment is to examine relevant scientific and technical information and develop conclusions based on multiple lines of evidence approach including consideration of persistence, exposure, trends, toxicity, bioaccumulation and widespread occurrence in the environment.

Assessment 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 type of organism of interest and dividing it by an assessment factor appropriate for the data point. A risk quotient (PEC/PNEC) is calculated for each type of endpoint in order to estimate potential ecological risk in Canada. However, there is also uncertainty regarding the risk that PFOA may pose now or in the future. Typically quantitative risk estimates (i.e., risk quotients or probabilistic analyses) are important lines of evidence when evaluating a substance’s potential to cause environmental harm. However, when risks for persistent and bioaccumulative substances such as PFOA are estimated using such quantitative methods, they are highly uncertain and are likely to be underestimated. Given that long term risks associated with persistent and bioaccumulative substances cannot at present be reliably predicted, quantitative risk estimates have limited relevance. Furthermore, since accumulations of such substances may be widespread and are difficult to reverse, a conservative response to uncertainty is justified.

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 assessment factor of 100 to give a PNEC of 0.158 µg/g or 158 µg/kg. The assessment factor was used to account for laboratory to field extrapolation, intra- and interspecies variation, the fact that the substance is persistent and has the potential to accumulate and biomagnify in terrestrial and marine mammals, 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.

Current Canadian surface water measurements include those of Etobico*ke 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 Etobico*ke Creek may be considered a high-density urban scenario for general surface waters in Canada owing to the population density surrounding Etobico*ke Creek,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 assessment factor of 100 to account for laboratory to field extrapolation, extrapolation from a LOAEL to a no-observed-adverse-effect level (NOAEL), 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 assessment factor of 100 to account for laboratory to field extrapolation, extrapolation from a LOAEL to a no-observed-adverse-effect level (NOAEL) 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.

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

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.

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Potential to Cause Harm to Human Health

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 in Lake Ontario and Lake Erie at levels ranging from 13 to 50 ng/L (mean of 8 samples from 4 locations in Lake Ontario: 42.5ng/L; mean of 8 samples from 4 locations in Lake Erie: 35.6 ng/L - Boulanger et al. 2004; median of 13 samples on Lake Ontario: 21 ng/L; median of 3 samples in 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 (Shoeib et al. 2011). 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; 8:2FTOH 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 8:2FTOH <LOD to 20 pg/m³ – 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. Neither PFOA 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 contain PFOA (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; Beesoon 2011), indicating potential exposures in utero and through breast milk.

The available data suggest that Canadians are exposed to PFOA 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 sources – 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 and migration from treated paper into food, and inhalation during the consumer use of spray treatment products on 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 blood 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, children 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 2–4 years (Olsen et al. 2007; Bartell at al. 2010; Brede et al. 2010).

PFOA has been measured in the blood of non-occupationally exposed adults, children, and newborns in several studies, at levels ranging from <0.0001 µg/mL up to 0.0561 µg/mL, presented in Table 7. Canadian biomonitoring data are consistent with data from the United States. Serum and plasma PFOA levels in children and the elderly are comparable with those in adults. However, based on data from a German study, infants aged 6 months appear to be the highest exposed age group. In the study, PFOA levels were measured at two time points in maternal blood (n~40 and 38), 6 months after delivery (n=47), cord blood (n=33) and in blood of infants six (n =40) and nineteen months (n=24) after birth, and monthly in breast milk samples. Concentrations in maternal serum ranged from 0.7 to 8.7 µg/L. In cord serum, the values ranged from 0.5 to 4.2 µg/L. The median results from serum at six and nineteen months of age were 6.9 and 4.6 µg/L, respectively. In breast milk, PFOA was detected only in some samples. Authors suggest the an increase in concentrations of PFOA during the first months of an infants life is due to breastfeeding, since breast milk contains measurable levels of PFOA (Fromme et al. 2010).

Most of the relevant toxicity studies in laboratory animals 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 2. 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.

A short-term (14 day) oral toxicity study with APFO in rats and mice identified a LOAEL of 0.3 mg/kg-bw per day. 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 serum PFOA 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). During the preparation of this final screening assessment, Macon et al. (2011) published a study investigating low dose effects of PFOA in mice. The study reported significantly lower mammary gland developmental scores in offspring of CD-1 mice on PND 21 when dams were exposed to 0.01 mg/kg-bw per day through oral gavage during GD 10-17; serum PFOA was 16.5±2.1 ng/ml. Authors reported a sensitivity of mammary glands to PFOA in CD-1 mice and that further research is needed to determine if the sensitivity is attributed to timing of exposure, mouse strain, or if there are other mechanisms underlying the apparent sensitivity in CD-1 mice. There is insufficient information to determine if these mammary gland effects are permanent and therefore adverse at low dose levels.

Two chronic toxicity studies in laboratory animals were identified. In the first chronic toxicity study, 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 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 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) (see Appendix 2 for specific compound details and references). 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 PPARα. Activation of PPARα causes 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 PPARα, rather than direct genotoxic action, is the critical event in the induction of these tumours. The confidence in the PPARα mode 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 PPARα in vitro and causes hepatic peroxisome proliferation in vivo. Temporal and dose–response data support a PPARα 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 β-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 PPARα activation 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 PPARα 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 PPARα 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 PPARα 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 PPARα agonist 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 to PFOA result from a PPARα-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 PPARα gene (PPARα-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 PPARα-null mice, suggesting that these effects are dependent on PPARα 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 PPARα-null mice. PFOA-exposed mice also had reduced thymus weight and number of thymocytes, effects that were attenuated in the PPARα-null mice. However, significant liver weight increase occurred in both wild-type and PPARα-null mice. Gene profiling studies also suggest that PFOA can alter mouse liver genes independent of PPARα (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. 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).

Several studies on occupationaly exposed workers (Olsen et al. 2003; Sakr et al. 2007a,b; Costa et al. 2009) and on highly exposed community residents (Steenland et al. 2009; Frisbee et al. 2010) have observed a positive association between serum PFOA and total cholesterol and other lipid parameters. Similar findings have been reported in a recent cross-sectional study on the general population of the United States (Nelson et al. 2010). In contrast, other occupational epidemiology studies (Olsen et al. 2000; Olsen and Zobel 2007) and highly exposed community studies (Emmett et al. 2006b) do not provide evidence of a significant association between serum PFOA and cholesterol.

The overall weight of evidence does not support causal relationships between PFOA and observed adverse health effects, due to numerous confounding factors including exposure to multiple chemicals. A recent review on the epidemiologic evidence of health effects of PFOA also suggests that the available epidemiological data are insufficient to make conclusions regarding the role of PFOA for any of the health effects observed (Steenland et al. 2010a).

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.

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 (0.3 mg/kg-bw per day) identified in critical studies, 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). In Macon et al. 2011, significantly lower mammary gland developmental scores in offspring of CD-1 mice on PND 21 were observed when dams were exposed to 0.01 mg/kg-bw/day. This study was not considered appropriate for use in risk characterization due to evidence of strain sensitivity, lack of repeatability of the endpoint at low exposure levels, lack of understanding of MOA, and the relevance of this endpoint for humans. Furthermore, the current study does not provide sufficient evidence that these effects are permanent and therefore adverse.

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 pharmaco*kinetics. 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 blood (serum or plasma) makes it the most appropriate measure of internal exposure. Due to the long half-life of PFOA and its lack of metabolism, human adult blood levels of PFOA represent cumulative (lifetime) exposure to PFOA and any contributions from its precursors. Blood 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, pharmaco*kinetic 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 PPARα in cell culture in vitro (Takacs and Abbott 2007; Wolf et al. 2008) but not male rat Leydig cells (Biegel et al. 2001). PPARα activation has a wide range of effects, including regulating the expression of genes involved in cell growth and survival. In fact, some PPARα 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), making PPARα a potential target for cancer therapy. It should also be noted that other PPARα 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.

The use of blood (serum or plasma) 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 blood provide a measure of aggregate exposure from multiple sources and exposure routes. The use of blood levels also significantly reduces uncertainties associated with interspecies and intraspecies differences in pharmaco*kinetics.

Use of 95th percentile serum or plasma 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). Although two older biomonitoring studies in the United States have shown declining serum PFOA levels from before the reduction measures were instituted (Calafat et al. 2007; Olsen et al. 2008), more recent data from NHANES demonstrate stable PFOA serum levels over the period from 2003 to 2008 (U.S. CDC 2011). Recent Canadian data exhibit a similar trend of stable levels over the period from 2004 to 2009 (see Table 7). The MOEs presented in Table 8 are based on serum and plasma levels from 1994 to 2009. PFOA levels in Canadians, as reported in the CHMS, are comparable to levels in the United States, as reported in the most recent NHANES data. There remains a gap in biomonitoring data for Canadian infants and children from birth to 2 years. Data from nursing infants have shown that this age group may represent the most highly exposed subpopulation (Fromme et al. 2010).

Uncertainty remains regarding the mode of action for tumour induction; however, as the database for genotoxicity suggests that PFOA 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.

There is some uncertainty regarding the effects of PFOA at low doses including mammary gland development; additional studies are required to further investigate these early findings.

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Conclusion

The assessment is based on a weight-of-evidence approach regarding the persistence, bioaccumulation, temporal trends in some species (i.e., the polar bear), long-range transport and the widespread occurrence and concentrations of PFOA in the environment and biota (including remote areas of Canada).

Based on the information presented in this screening assessment, it is concluded that PFOA, its salts and its precursors are entering or may enter 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. In addition, it is concluded that PFOA and its salts are extremely persistent and meet the criteria for persistence as set out in the Persistence and Bioaccumulation Regulations. PFOA and its salts do not meet the criteria for bioaccumulation as set out in the Persistence and Bioaccumulation Regulations. Nevertheless, the weight of evidence is sufficient to conclude that PFOA and its salts accumulate and biomagnify in terrestrial and marine mammals

Based on a comparison of upper-bound levels of PFOA in the blood (plasma or 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 (Table 8) are adequately protective of human health. It is concluded 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 that can degrade to PFOA in the environment were not individually assessed. The approach for the assessment included the contribution of precursors to total PFOA which is the moiety of toxicological concern, as they can degrade to PFOA in the environment.

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

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Appendix 1: Summary of health effects information for PFOA