Draft Ecological Screening Assessment Report Long-Chain (C9–C20) Perfluorocarboxylic Acids, their Salts and their Precursors

The information below was considered in evaluating whether long-chain (C9-C20) perfluorocarboxylic acids, their salts and their precursors meet the criteria for persistence and bioaccumulation as defined under the Persistence and Bioaccumulation Regulations (Canada 2000). Persistence criteria are half-lives of greater than or equal to 2 days in air, 182 days in water, 365 days in sediment, or 182 days in soil, or evidence of long-range transport to remote areas. Bioaccumulation criteria are bioaccumulation factors (BAFs) or bioconcentration factors (BCFs) of greater than or equal to 5000 or a log K_OW of greater than or equal to 5.0.

Persistence

A small amount of C9 PFCA was produced through a photo-induced hydrogen peroxide system, demonstrating rapid degradation of 10 and 100 µg/M solutions of 8:2 FTOH within minutes to hours via the formation of 8:2 FTAL (fluorotelomer aldehyde), 8:2 FTCA and 8:2 FTUCA (Gauthier and Mabury 2005). However, this is an aqueous photolysis reaction that may not significantly contribute to long-chain (C9-C20) PFCAs in the environment given the low aqueous solubility and large Henry’s Law constant of FTOHs.

Hori et al. (2005a) have reported C9 PFCA decomposition where the concentration of C9 was 1.51 mg/L. Hori et al (2005b) also examined the degradation of C9, C10 and C11 PFCA with persulfate ion (S₂O₈ ^-2) in an aqueous/liquid CO₂ biphasic system. C9 PFCA was degraded to fluoride ions and carbon dioxide in a solution containing S₂O₈^-2 heated to 80°C for 6 hours (Hori et al. 2008). However, the conditions in these studies are not environmentally relevant.

Hurley et al. (2004) have shown that atmospheric degradation lifetime of gas phase short-chain PFCAs (C3–C5), under artificial smog conditions, is expected to be on the order of 130 days due to OH radical reactions, with a lifetime of the order of 10 days due to wet/dry deposition (particle mediated). Direct gas phase photolysis of the acids was not observed. Hurley et al. (2004) also stated that it is unlikely that these values will significantly change as the chain length of the acid is increased. The degradation pathway initiated by the reaction C_nF_2n+1COOH + OH ® H₂O + C_nF_2n+1COO (followed by C_nF_2n+1COO ® C_nF_2n+1 + CO₂, etc.) is not believed to be particularly efficient, given that the lifetime for this process (130 d) is considerably greater than that estimated for removal of PFCAs from the atmosphere by wet/dry deposition (~ 10 d). In other words, even if PFCAs are formed in the atmosphere from FTOHs, they will not remain there long enough to be degraded.

The presence of long-chain PFCAs in the Canadian Arctic (Martin et al. 2004a) indicates the long-range transport either of long-chain (C9-C20) PFCAs (e.g., via ocean currents) (Wania 2007; Prevedouros et al. 2006) or of volatile precursors to long-chain (C9-C20) PFCAs such as FTOHs (e.g., via atmospheric transport) or both (Wallington et al. 2006, Stock et al. 2007). C9 to C11 PFCAs were measured in polar ice caps from three areas in the High Arctic in the spring of 2005 and 2006 (Melville ice cap, Northwest Territories; Agassiz ice cap, Nunavut; and Devon ice cap, Nunavut) (Young et al. 2007). C9 PFCA concentrations ranged from 0.005 to 0.246 ng/L. C10 PFCA concentrations ranged from below detection to 0.022 ng/L. C11 PFCA concentrations ranged from below detection to 0.027 ng/L. Between 1996 and 2005, concentrations were increasing for C9 and C10 PFCAs (Young et al. 2007). Fluxes were calculated using the density corrected concentration, multiplied by the yearly accumulation. Fluxes calculated to each of the ice caps were multiplied by the area of the Arctic to yield a flux of C9, C10, and C11 PFCAs to the area north of 65°N. These fluxes are estimates and may not be representative of actual deposition in this region due to wide variations in precipitation rates. In 2005, C9 showed a flux ranging from 73- 860 kg/year; C10 PFCA showed a flux ranging from 16 – 84 kg/year and C11 PFCA showed a flux ranging from 26-62 kg/year (Young et al. 2007).

A suggested hypothesis for the presence of long-chain (C9-C20) PFCAs in biota in remote regions is that a precursor (e.g., FTOHs) is emitted to the atmosphere and ultimately degrades to yield long-chain (C9-C20) PFCAs through biotic and abiotic degradation. Ellis et al. (2004a) showed that the atmospheric lifetime of short-chain FTOHs, as determined by their reaction with hydroxy radicals, was approximately 20 days. Shoeib et al. (2006) collected air samples during a crossing of the North Atlantic and Canadian Archipelago in July 2005 to investigate concentrations of FTOHs. The highest concentrations were for 8:2 FTOH at 5.8–26 pg/m³, followed by 10:2 FTOH at 1.9–17 pg/m³and then 6:2 FTOH at below detection limit to 6.0 pg/m³. The surfactant properties of PFCAs have been examined for their influence on the potential formation of perfluorinated aerosols over the marine environment (Waterland et al. 2005) and may suggest a mechanism for long-range transport to remote regions via oceanic routes. However, available research suggests that the presence of long-chain (C9-C20) PFCAs in remote regions may be a result of the degradation of volatile fluoroalkyl precursor substances such as FTOHs. Young et al. (2007) suggested that the presence of C9, C10 and C11 PFCAs on Canadian High Arctic ice caps is indicative of atmospheric oxidation of volatile precursors as a source. High Arctic ice caps receive contamination from atmospheric sources and have the potential to provide long-term temporal trends of atmospheric concentrations. The ratios of these PFCAs to sodium concentrations were variable, indicating that PFCAconcentrations on the ice cap are unrelated to marine chemistry. By examining the concentrations of PFCAs in ice caps, atmospheric fluxes were determined by considering the area of each ice cap. Fluxes of PFCAs were estimated for the area north of 65°N for 2005 to be 73–860 kg/year for C9, 16–84 kg/year for C10 and 26–62 kg/year for C11.

The carbon-fluorine bond is one of the strongest in nature (~110 kcal/mol), making the bond extremely stable and generally resistant to degradation. Fluorine has the highest electronegativity of all elements in the periodic table. This contributes to a high ionization potential and low polarizabilty. It also results in low inter- and intra-molecular interactions and extremely low surface tension. Direct photolysis of a carbon-fluorine chain is also expected to be very slow, with stability to such energy expected to be sustained for more than 1000 years (Environment Canada and Health Canada 2006). PFCAs, in general, have surfactant properties due to the combined properties of oleophobicity, hydrophobicity, and hydrophilicity over portions of a particular molecule. The nature of PFCAs, e.g., their strong tendency to ionize, would likely lead to them to be more prevalent in the aqueous phase, and they are not expected to partition significantly to the atmosphere (Ellis et al. 2004a). It is unlikely that these acids will degrade under environmentally relevant conditions in any medium, including water and/or sediment.

Bioaccumulation

Regulatory criteria (BCFs and BAFs) have been developed under the Canadian Environmental Protection Act, 1999 (Canada 1999) to determine whether or not 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 1995). These criteria were intended to identify lipophilic substances with the potential to bioaccumulate primarily in aquatic systems.Thus, 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. It should be noted, however, that information on BAFs, BCFs or log K_ows is but one part of the overall weight of evidence in determining the bioaccumulation potential of a given substance. Furthermore, a substance may be deemed to be sufficiently bioaccumulative to cause concern, even if regulatory criteria are not met.

Additional measures of bioaccumulation which directly address the potential for chemicals to biomagnify include biomagnification factors (BMFs) and trophic magnification factors (TMFs; sometimes referred to as food-web biomagnification factors). The BMF represents the ratio of the chemical concentration in a predator to that in its food or prey. A BMF greater than 1 may be considered cause for concern, as it suggests that biomagnification is occurring. The BMF measured relative to a food item in the laboratory is sometimes referred to as a “dietary BAF.” An important uncertainty in BMF measurements is associated with determining the actual trophic status of a predator and its prey, given that most organisms are omnivores (Gray 2002). A TMF may be thought of as the average ratio of the concentration of a substance in predator and prey across an entire or partial food web. As with BMF, a TMF value exceeding 1 may also be cause for concern, since it indicates that food web biomagnification is occurring. BMFs and TMFs are most often measured in the field, although laboratory feeding studies can also be used to estimate BMFs (or “dietary BAFs”). Generally, chemical concentrations are lipid-normalized prior to making BMF and TMFdeterminations; however, lipid-normalizing concentrations of perfluorinated substances may not be appropriate, since these substances appear to preferentially bind to proteins in liver, kidney and plasma rather than partition to lipids (Houde et al. 2006b; Martin et al. 2003a). The lack of a normalization method for substances that associate with protein/plasma introduces a source of uncertainty when evaluating BMFs and TMFs of PFCAs.

PFCAs have the combined properties of oleophobicity, hydrophobicity, and hydrophilicity over different portions of these molecules. Furthermore, the carboxylate functional group attached to the perfluorinated chain imparts polarity to the molecule. Due to these properties, the normal assumption that the hydrophobic and lipophilic interactions between compound and substrate are the main mechanisms governing partitioning may not be applicable for long-chain (C9-C20) PFCAs. Long-chain (C9-C20) PFCAs are considered to be “model-difficult” (i.e., most models use log K_ow which is not applicable to perfluorinated substances due to their partitioning preference to proteins) with respect to bioaccumulation, and standard measures of bioaccumulation should be applied cautiously for these substances. There are no reported K_ow measurements for any long-chain (C9-C20) PFCA, and the use of this physical property for estimation of bioaccumulation potential is unlikely to be useful because these substances can sit at the interphase between organic and aqueous phases rather than partition between the two phases (Houde et al. 2006b).

It has been suggested that an additional assumption of the BAF/BCF/log K_ow approach is that bioaccumulation occurs by the same mechanisms for all chemicals in both water-breathing animals (e.g., fish and aquatic invertebrates) and air-breathing animals (e.g., terrestrial mammals, birds and marine mammals), resulting in a similar bioaccumulation potential between these organism classes for a particular substance (Kelly et al. 2004; Mackay and Fraser 2000). As described by Kelly et al. (2004), organic chemicals can be grouped according to polarity (as indicated by a log K_owthat decreases with increasing polarity due to expected changes in aqueous solubility), and volatility (as indicated by a log K_oa (octanol-air partition coefficient) that decreases with increasing volatility). In general, non-polar, non-volatile (NPNV) chemicals such as PCBs are expected to have low elimination rates to both water and air, resulting in a similarly high bioaccumulation potential for both air-breathing and water-breathing organisms. The polar nature of polar, non-volatile (PNV) chemicals and the potential ionization of PFCAs in particular, will cause their water solubility to increase relative to NPNVs. For water-breathing organisms, this potentially results in more rapid elimination of PNVs to the water phase and a reduction in bioaccumulation potential. However, because bioaccumulation potential in air-breathing organisms is driven primarily by volatility rather than polarity, the non-volatile nature of PNVs such as PFCAs results in their relatively slow elimination to air, resulting in higher bioaccumulation potential in air breathers (Stevenson 2006b).

Although the general assumption is that chemical properties and partitioning behaviour are the primary processes governing uptake and elimination, in many cases metabolic transformation of a particular chemical allows for rapid elimination and lower bioaccumulation potential (Kelly et al. 2004). However, studies have not been performed on the metabolic transformation and elimination of PFCAs or precursors in air-breathing organisms.

An additional complication relating to bioaccumulation assessment for long-chain (C9-C20) PFCAs is that BCFs, BAFs, BMFs and TMFs are often based on concentrations in individual organs, as opposed to whole-body burdens. From a toxicological perspective, BCFs, BAFs, BMFs and TMFs for individual organs, such as the liver, may be more relevant when predicting potential for direct organ-specific toxicity (e.g, liver toxicity). Conder et al. (2008) suggest that, as bioaccumulation is expressed on a whole-body mass–basis, the concentrations of perfluorinated acids in tissues such as liver are not appropriate for use in assessing the bioaccumulation potential of these compounds. Due to the small proportion of the body mass that is composed of liver tissue and blood and the magnitude of the differences in concentrations between these compartments and other tissues, the concentration of perfluorinated acids on a whole-body mass–basis has been estimated to be 10 times lower than concentrations of perfluorinated acids in plasma in dolphins, narwhal and beluga whale and 2–10 times lower than the concentrations of perfluorinated acids in blood and liver of trout (Conder et al. 2008).

However, measures of bioaccumulation (BCFs, BAFs, BMFs) may be used as indicators of either direct toxicity to organisms that have accumulated long-chain (C9-C20) PFCAs or of indirect toxicity to organisms that consume prey containing long-chain (C9-C20) PFCAs (via food chain transfer). Concerning the potential to cause direct toxicity, the critical body burden is the minimum concentration of a substance in an organism that causes an adverse effect. From a physiological perspective, it is the concentration of a substance at the site of toxic action within the organism that determines whether a response is observed, regardless of the external concentration. In the case of long-chain (C9-C20) PFCAs, the site of toxic action is often considered to be the liver. Concerning the potential for toxicity to consumer organisms, it is the concentration in the whole body of a prey 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. Given the partitioning into liver and blood, most field measurements for perfluorinated substances have been performed for those individual organs and tissues especially for higher-trophic-level organisms (e.g., polar bear) where whole-body analysis is not feasible due to either sampling or laboratory processing constraints. While it is feasible to measure whole-body BAFs on smaller, lower-trophic-level species, the lower trophic status of the organism would mean that, for perfluorinated substances, the estimated overall BAFs may be underestimated due to their trophic status. Thus, from a toxicological perspective, BCFs, BAFs and BMFs based on concentrations in individual organs, such as the liver, may be more relevant when predicting potential for direct organ-specific toxicity (i.e., liver toxicity). BCFs and particularly BMFs based on concentrations in whole organisms may provide a useful measure of overall potential for food chain transfer. Conder et al. (2008) suggested that BMF values are relevant for bioaccumulation potential in higher-trophic-level biota, as extrapolating BCF/BAF data for fish and invertebrates is difficult due to the biological differences between the higher and lower trophic levels.

Bioaccumulation/Bioconcentration/Biomagnification Studies

Bioaccumulation of C9 to C12 PFCAs from laboratory-spiked and contaminated field sediments was assessed using the freshwater oligochaete, Lumbriculus variegatus, a deposit feeder that can serve as an entry point for sediment-bound contaminants into food webs (Higgins et al. 2007). Semi-static batch experiments were conducted over 56 days. It should be noted that the sediment concentrations in the laboratory-spiked systems decreased slightly over time, whereas the sediment concentrations for nearly all the long-chain PFCAs in the contaminated field sediment remained essentially constant. The biota-sediment accumulation factors (BSAF), wet weight (ww), were as follows: C9 (0.64–1.60), C10 (0.59–1.02), C11 (0.42–0.62) and C12 (0.42–0.55). The authors suggest that the long-chain PFCAs may not have reached steady-state.

Martin et al. (2003a,b) used juvenile rainbow trout (Oncorhynchus mykiss), dietary exposure and a flow-through aqueous exposure using C9-C14 PFCAs. BCFs for rainbow trout increased as perfluoroalkyl chain lengths increased, with reported whole-body values from 450 L/kg for C10 PFCA to 23 000 L/kg for C14 PFCA (Martin et al. 2003b). No experimental data were available for C9 PFCA because it was used as an internal standard in these studies. For the juvenile rainbow trout dietary exposure study, Martin et al. (2003a,b) also report “dietary BAFs.” However, based on an examination of the accumulation equation and given that exposure was via the diet rather than water, it can be concluded that the measurements were actually equivalent to BMFs. The lab-measured BMFs for rainbow trout showed an increasing trend approaching 1 for C14 PFCA. The authors speculated that the lack of observed biomagnification (i.e., no BMFs exceeded 1) was likely due to the small size of fish used in the study, resulting in more rapid chemical elimination to water, relative to body size, than would be observed for larger species or size classes. This more rapid chemical elimination would reduce the BMF.

Martin et al. (2004b) also conducted a field study of the biomagnification of C9 to C14 PFCAs in the pelagic food web of Lake Ontario and determined lake trout (Salvelinus namaycush) BMFs for a variety of prey species (alewife – Alosa pseudoharengus; rainbow smelt – Osmerus mordax; and slimy sculpin – Cottus cognatus), as well as overall TMFs for the pelagic food web. Lake trout / alewife BMFs exceeded 1 for all long-chain PFCAs measured in the study (C9-C14); lake trout / smelt BMFs ranged from 0.6 (C9) to 2.2 (C14); and lake trout/sculpin BMFs ranged from 0.1 (C9) to 0.4 (C13). The authors report that alewife comprise 90% of lake trout prey, suggesting that lake trout/alewife results provide the best BMFestimates. Given that the other prey species comprised a much lower proportion of the diet of lake trout (7% for smelt and 2% for sculpin), the lake trout BMF estimates for these prey are likely to be less reliable. In particular, the authors cautioned that the low dietary proportion of sculpin for lake trout and the position of sculpin in the benthic rather than pelagic food web could explain the low BMFs observed for lake trout/sculpin. To address differences in dietary composition, the authors also calculated lake trout/prey BMFs that weighted the concentration in each prey species with the proportion of each prey species in the diet. The resulting BMFs were above 1 for all of the C9–C14 PFCAs, indicating biomagnification from consumed prey for the lake trout of Lake Ontario.

Trophic magnification factors (TMFs) measured in the pelagic aquatic food web of Lake Ontario by Martin et al. (2004b) suggest trophic magnification for some long-chain PFCAs over the whole food web. Concentrations of C10, C11 and C13 PFCAs increased significantly within the pelagic food web, resulting in TMFs greater than 1 for C10, C11 and C13. Trophic magnification was greatest for C11 PFCA (4.7) and decreased for longer and shorter PFCAs alike. TMFs equal to 1 for C9, C12 and C14 PFCAs indicated either no biomagnification or that the results were too variable to detect a statistically significant trend in concentration with trophic status for this food web.

Gulkowska et al. (2005) analyzed avian and fish blood samples and water samples from the Gulf of Gdansk for C9 PFCA. Sixty-five blood samples were collected during winter 2002–2003 from five species of waterfowl--razorbill (Alca torda), red-throated loon (Gavia stellata), black scoter (Melanitta nigra), long-tailed duck (Clangula hyemalis) and common eider (Somateria molissima)--while 18 blood samples were collected from cod (Gadus morhua). The mean concentration of C9 PFCA in avian blood samples ranged from 0.3 ng/ml in razorbill to 1.1 ng/ml in red-throated loon. The mean concentration of C9 PFCA in cod blood samples was 1.2 ng/ml. The authors reported a blood:water “BCF” for C9 PFCA in cod of approximately 3000. However, given that this measurement was field-based, where the cod would be exposed via water and the diet, the reported BCF is analogous to a BAF. The bird/cod BMFs ranged from 0.25 to 0.92, but the authors cautioned that all bird species sampled were migratory and it is unclear whether they included a large proportion of cod in their diet. There is also uncertainty as to whether the blood-based BMFs would be similar to whole-body BMFs.

HaukÅs et al. (2007) determined C9 PFCA BMFs for a Barents Sea (east of Svalbard) ice edge food web comprising of the ice-associated amphipod (Gammarus wilkitzkii), polar cod (Boreogadus saida), black guillemot (Cepphus grylle), and glaucous gull (Larus hyperboreus). BMFs were not calculated for the amphipod as C9 PFCA was not quantifiable. However, the BMFfor the black guillemot/polar cod was calculated to be 8.76; the BMF for the glaucous gull/polar cod was 11.6; and the BMF for the glaucous gull/black guillemot was 9.34.

Tomy et al. (2009c) determined TMFs for C9 – C11 PFCAs 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 (C10, Arctic cod/Themisto libellula) to 353 (C11, beluga whale/Pacific herring).

Houde et al. (2006a) conducted field studies of the bottlenose dolphin food web in Charleston, South Carolina, and Sarasota Bay , Florida. C9 to C12 PFCAs were measured in seawater, marine sediment, zooplankton (Sarasota Bay only; species not identified) and a variety of fish species: Atlantic croaker (Micropogonias undulates), pinfish (Lagodon rhomboids), red drum (Sciaenops ocellatus), spotfish (Leiostomus xanthurus), spotted seatrout (Cynoscion nebulosus), striped mullet (Mugil cephalus) and bottlenose dolphin (Tursiops truncatus). It should be noted that for this particular study, samples were collected over a series of years, and prey and predator species may have been collected in different years/seasons, which may impact the BMFs and TMFs reported. Fish were captured from 2002 to 2004. Zooplankton samples were collected in 2004. Dolphin plasma, skin and teeth were collected from both locations in summer 2004. Recently deceased bottlenose dolphins from 2002 and 2003 were also used. Dolphin samples included plasma from a catch-and-release study and multiple whole-body samples from recently deceased or stranded dolphins, facilitating an examination of trophic magnification in terms of both dolphin plasma and whole body. BMFs were reported for whole-body concentrations only. For Charleston, marine fish BMFs (sea trout/pinfish) ranged from 0.1 (C12) to 3.7 (C10), with no clear trend with chain length. Dolphin BMFs were reported for whole-body samples and a wide range of prey fish species. BMFs exceeded 1 for all dolphin/prey combinations for C9, C10 and C11 PFCAs. For C12 PFCA, the dolphin/prey BMFs ranged from 0.1 to 1.8. In Sarasota Bay, BMFs were reported for C12 PFCA only and ranged from 0.2 to 156 for fish/prey (multiple species) and measured 0.1 for dolphin/striped mullet. TMFs for the dolphin food web were only reported for Charleston. For C9–11 PFCAs, both whole-body- and plasma-based TMFs exceeded 1, while for C12 PFCAneither the plasma nor the whole-body TMF exceeded 1. Dolphin BMFs exceeding 1 for C9–C11 PFCAs suggest that these PFCAs are biomagnifying from fish to dolphins in this food web. For C12 PFCA, the range in BMFs makes it difficult to draw conclusions regarding biomagnification without knowledge of the feeding preferences of bottlenose dolphins. The evidence for fish-to-fish biomagnification is mixed; however, biomagnification might be expected to be lower in fish than in dolphins given that PFCAs may be eliminated more rapidly to water than to air. The TMF results integrate the findings for the whole food web. Despite the expected lower biomagnification potential in fish, TMFs for C9–C11 PFCAs exceeded 1 for the dolphin food web, indicating that trophic magnification is occurring.

van den Heuvel-Greve et al. (2009) determined C11 PFCA BMFs in the harbour seal (Phoca vitulina) food web in the Westerschelde, an estuary in the southwest of the Netherlands. The BMFs ranged from 1.9 (herring:zooplankton) to 53 (harbour seal:herring) with a TMF of 1.3.

Katz et al. (2009) showed that C9-C12 PFCAs were bioaccumulating in the vegetation (plants and lichens)-barren ground caribou (Rangifer tarandus groenlandicus)-wolf (Canis lupus) terrestrial food chain in northern Yukon, Canada. Lichens reflect direct atmospheric input of long-chain PFCAs as they lack roots and receive their nutrients from the atmosphere. Lichen is large part of the caribou diet. Caribou are the staple prey of wolves – the top predator in the ecosystem. C9 PFCA was dominant in wolf liver at 6.8 ng/g ww followed by C10 PFCA at 3.1 ng/g ww and C11 PFCA at 3.4 ng/g ww. C12 and C13 PFCAs were also measured with average concentrations < 0.6 ng/g ww. BAFs or BMFs were not calculated. However, the results of the carbon and nitrogen stable isotope analyses of the vegetation, caribou muscle and wolf muscle showed that the caribou were primarily feeding on the lichen and that the wolves were feeding primarily on the caribou.

Powley et al. (2008) determined C10-C12 PFCA bioaccumulation factors for a western Canadian Arctic (Banks Island on the eastern edge of the Beaufort Sea in the Northwest Territories) food web comprising of three different species of zooplankton (Calanis hyperboreus, Themisto libellula, and Chaetognatha), Arctic cod (Boreogadus saida), ringed seal (Phoca hispida), and bearded seal (Eriganthus barbatus). C11 PFCA had the highest concentration at 10.8 ng/g. Bioaccumulation factors ranged from 0.3 to 3.1.

Multiple investigations (Martin et al. 2004a; Kannan et al. 2005; Smithwick et al. 2005a) have also found concentrations of C9 (108–230 ng/g-ww), C10 (35–76 ng/kg-ww), C11 (56–78 ng/g-ww), C12 (4.7–8.2 ng/kg^-ww),C13 (7.5–14 ng/g-ww), and C14 (< 0.5–1.1 ng/kg-ww) in polar bear livers located in the Canadian Arctic and sub-Arctic regions. Butt et al. (2008) calculated regionally-based for ringed seal liver –polar bear liver BMFs for C9 to C15 PFCAs by grouping 11 populations of ringed seals to correspondingly, similarly located polar bear populations. BMF geometric means ranged from 2.2 (C13 PFCA) to 56 (C9 PFCA).

There are no bioaccumulation studies for long-chain PFCAs greater than C14. However, there is the potential that long-chain PFCAs greater than C14 could bioaccumulate or biomagnify in marine and/or terrestrial species. It has been suggested the carbon-carbon conformation changes as the chain length increases, with longer chains becoming helical (Wang and Ober 1999), resulting in smaller cross-sectional diameter molecules with the ability to accumulate in organisms. C14 and C15 have been found in fish, invertebrates, dolphin and polar bears (e.g., Martin et al. 2004b; Smithwick et al. 2005a, 2005b, 2006; Houde et al. 2005). Conder et al. (2008) highlighted, based on available research on the bioaccumulation of perfluorinated acids, five key points:

A) Bioconcentration and bioaccumulation of perfluorinated acids are directly related to the length of each compound’s fluorinated carbon chain

B) Perfluorosulfonic acids are more bioaccumulative than PFCAs of the same fluorinated carbon chain length.

C) PFCAs with seven fluorinated carbons or less (perfluorooctanoate (PFO) and shorter PFCAs) are not considered bioaccumulative according to the range of international promulgated bioaccumulation regulatory criteria of 1000–5000 L/kg.

D) PFCAs with seven fluorinated carbons or less have low biomagnification potential in food webs.

E) More research is necessary to fully characterize the bioaccumulation potential of PFCAs with longer fluorinated carbon chains (> 7 fluorinated carbons), as PFCAs with longer fluorinated carbon chains may exhibit partitioning behavior similar to or greater than PFOS.

Table 3. Summary of Bioaccumulation Data for Long-Chain (C9-C20) PFCAs