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

Direct Releases

There are no available data on the direct release of long-chain (C9-C20) PFCAs to the Canadian environment.

Indirect Releases

There is empirical evidence available regarding the degradation of fluorotelomer-based polymers into long-chain (C9-C20) PFCAs. Fluorotelomer alcohols (FTOHs) with x number of carbons produces intermediates such as fluorotelomer unsaturated carboxylates (x:2 FTUCA) where x equals the number of carbons and fluorotelomer carboxylic acids( x:2 FTCA) that can further degrade to a long-chain (C9-C20) PFCAs. FTOHs can be biodegraded or metabolized to long-chain (C9-C20) PFCAs as shown in studies by Hagenet al. 1981, Lange 2002, Dinglasanet al. 2004, Kudoet al. 2005, Martinet al. 2005, Wanget al. 2005a, 2005b; Fasanoet al. 2006, 2008; Liuet al. 2007, and Nabbet al. 2007. Further evidence that FTUCAs and FTCAs are formed as intermediates in the biodegradation or metabolism of FTOHs is provided by Kudo et al. (2005), Martinet al. (2005) and Liuet al. (2007).

Recognition of FTOHs as potential sources of long-chain (C9-C20) PFCAs came from the detection of FTOH metabolites in biota (Smithwick et al. 2006; Butt et al. 2008; Powley et al. 2008; Furdui et al. 2007). Metabolism of FTOHs is expected to result in the formation of intermediates such as FTCAs and FTUCAs (Dinglasan et al. 2004; Wang et al. 2005a, 2005b). Houde et al. (2005) reported levels of 8:2 and 10:2 FTUCAs in plasma of bottlenose dolphins sampled from the region of the Gulf of Mexico along the Eastern coast of the Atlantic. FTCAs were not detected. A temporal trend study by Butt et al. (2008) also reported levels of FTUCAs in all ringed seal liver samples from the Canadian Arctic. Furdui et al. (2007) reported 8:2 FTUCA and 10:2 FTUCA in 52% and in 40% of all samples of lake trout from the Great Lakes, respectively. The presence of FTCAs and FTUCAs in animal biota is also reported in a number of studies such as Taniyasu et al. 2005, Verreaultet al. 2005, Powleyet al. 2008, Smithwick et al. 2006, Buttet al. 2007a, 2007b, 2008; and Furdui et al. 2007. Dinglasan and Mabury (2005) showed that 8:2 FTOH, 8:2 FTCA and 8:2 FTUCA are formed through the degradation of an 8:2 telomer methacrylate monomer that is used in building polymers. Although the rate of degradation was not determined, the aerobic sewage treatment plant innoculum was able to significantly degrade the monomer over the ~73-day test.

The relative abundance of linear vs. branched forms of PFCAs can provide some indication of their potential source. For example, DeSilva and Mabury (2004) showed that liver samples had at least 99% linear PFCA isomers in polar bears from southeastern Hudson Bay and eastern Greenland. Linear PFCAs reflect degradation largely from linear FTOHs and indicate that the source of PFCAs originates from telomerization rather than electrochemical fluorination (a process that would produce about 20% branched isomers). Additional indications for FTOHs as a source of PFCAs come from the odd–chain length and even–chain length patterns from PFCAs detected in tissue samples. It can be expected that the degradation of a given FTOH would result in the formation of an equal number of adjacent odd–chain length and even–chain length PFCAs via atmospheric oxidation. FTOH has been shown to degrade to a relatively equal concentration of even–chain length and odd–chain length PFCAs (Ellis et al. 2004b) and therefore, the exposure to each should be the same in lower-trophic-level biota. The odd–chain length PFCAs would be expected to be found at slightly higher levels in higher-trophic-level biota (Martin et al. 2004a). Such a pattern is seen in observations in polar bear sampling by Kannan et al. (2005) and Smithwick et al. (2005b). The correlated odd-chain length and even-chain length PFCAs tend to indicate a single uniform source (Smithwick et al. 2005b; van de Vijver et al. 2005) and could, therefore, be reflective of the manufacture of fluorotelomer alcohols. FTOHs appear to be available to biota in the environment and are being metabolized, in vivo, to intermediates (other PFCA precursors) which may ultimately yield long-chain PFCAs. Furdui et al. (2008) detected branched C11 PFCA and C13 PFCA isomers in lake trout from Lake Ontario that declined from 1993 to 2004 and then linear isomers increased in more recent samples (up to year 2004) suggesting that current PFCA sources to Lake Ontario result from the telomerization process.

The levels of residual FTOHs in polymers were measured in a study by Dinglasan-Panlilio and Mabury (2006) where several products containing fluorinated polymers or related fluorochemicals were analyzed. FTOHs (4:2 to 12:2) were found in products at levels between 0.11 and 3.8% on a dry weight basis. Extraction solvent was ethyl acetate of 2 x 5 ml aliquots which were subsequently combined. The concentration of the ethyl acetate was not provided. Although the actual levels of FTOHs present as residual^[1] or present as part of product formulations could not be distinguished, their presence provides some indication that fluorotelomer-based polymers could be a source of FTOHs to the environment.

The levels of long-chain (C9-C20) PFCAs measured in Canadian urban aquatic compartments suggest indirect input sources, e.g., wastewater treatment plants (WWTPs) (Boulanger et al. 2005a; Simcik and Dorweiler 2005; Crozier et al. 2005). C9 to C12 PFCAs have been detected in WWTP sludge in a number of studies (Boulanger et al. 2005b; Higgins et al. 2005; Sinclair and Kannan 2006; Crozier et al. 2005). Higgins et al. (2005) indicated higher levels of even-chained-length PFCAs (C8 to C12) in aerobically digested sludge from a WWTP and in sediment from the San Francisco Bay area. Sinclair and Kannan (2006) reported a pattern of higher even-chained-length PFCAs over odd-chained-length PFCAs in WWTP effluent waters in plants in New York State.

WWTPs with simple primary treatment did not have releases of long-chain (C9-C20) PFCAs. However, WWTPs that included secondary treatment increased the presence of long-chain (C9-C20) PFCAs (Sinclair and Kannan 2006), suggesting rapid biological or chemical degradation of precursors during secondary treatment. 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). As the FTCAs are only found in primary treatment samples, this suggests that the conversion of FTOHs to long-chain (C9-C20) PFCAs is incomplete, whereas the absence of FTCAs and presence of C9 to C11 PFCAs in secondary samples suggests complete conversion. Crozier et al. (2005) measured levels of C9 and C10 PFCAs in effluent waters (concentrations ranging from 3 – 6 ng/L) and biosolids (concentrations ranging from 0.4 – 5.2 ng/g) from Ontario sewage treatment plants. C11 and C12 PFCAs were not detected (detection limit 2 ng/L). Crozier et al. (2005) also noted that C10 PFCA which was not detected in the influent of one sewage treatment plant but was detected in the effluent at 4 ng/L indicating that C10 PFCA was formed during the sewage treatment plant process whereas C9 PFCA was detected at 4 ng/L in both the influent and effluent of the sewage treatment plant, suggesting no removal of the compound.

De Silva et al. (2009) suggested that the biodegradation and/or metabolism of polyfluorophosphoric acids (diesters equals diPAPs) such as 10:2 diPAP can yield C10 PFCAs. diPAPs were measured at 50 -100 ng/g in wastewater treatment plant (WWTP) sludge.

Guo et al. (2009) detected C9 to C12 PFCAs in typical American homes with carpeted floors, pre-treated carpet, and commercial carpet-care liquids. Floor waxes and stone/tile/wood sealants that contain fluorotelomer products are potential sources of C9 to C12 PFCAs in homes containing these materials. Other potential sources include treated home textile, upholstery and apparel and household carpet/fabric care liquids and foams (Guo et al. 2009). The release of PFCA precursors from household products is shown by several studies in which indoor air in houses was sampled. Archived U.S. house dust samples collected between 2000 and 2001 from Ohio and North Carolina were analyzed for FTOHs (6:2, 8:2 and 10:2) and PFCAs (C9–C12) (Strynar and Lindstrom 2005). Mean concentrations were 0.5–0.804 µg/g of dust for C9–C12 PFCAs. Mean 6:2, 8:2, and 10:2 FTOH levels ranged from 0.4 to 1.0 µg/g dust. The mean values between the two locations did not differ significantly, suggesting similar sources such as treated carpets or textiles. Shoeib et al. (2005) reported levels of 6:2, 8:2 and 10:2 FTOH in indoor dust collected from vacuum cleaners from randomly selected homes in Ottawa, with mean concentrations of 0.035, 0.055 and 0.035 µg/g of dust, respectively. Air samples for FTOH analysis were not collected due to technical difficulties. FTOHs have also been found in all-weather clothing (Berger and Herzke 2006) and as emissions from non-stick frying pans (Sinclairet al. 2007). PFCAs themselves (and in some cases, FTCAs and FTUCAs) may also be released from products, including all-weather clothing, cookware, commercial fabric protector and food contact materials (Begleyet al. 2005; Boulangeret al. 2005b; Mawnet al. 2005; Washburnet al. 2005; Bradleyet al. 2007; Sinclairet al. 2007).

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