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.

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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.