Environ. Sci. Technol. 2007, 41, 3521-3528
Prevalence of Long-Chained Perfluorinated Carboxylates in Seabirds from the Canadian Arctic between 1975 and 2004 CRAIG M. BUTT,† SCOTT A. MABURY,† DEREK C.G. MUIR,‡ AND B I R G I T M . B R A U N E * ,§ Department of Chemistry, University of Toronto, 80 St. George Street, Toronto, Ontario M5S 3H6, Canada, Water Science and Technology Branch, Environment Canada, 867 Lakeshore Road, Burlington, Ontario L7R 4A6, Canada, and National Wildlife Research Centre, Environment Canada, Carleton University, Raven Road, Ottawa, Ontario K1A 0H3, Canada
Temporal trends in perfluoroalkyl compounds (PFCs) were investigated in liver samples from two seabird species, thickbilled murres (Uria lomvia) and northern fulmars (Fulmaris glacialis), from Prince Leopold Island in the Canadian Arctic. Thick-billed murre samples were from 1975, 1993, and 2004, whereas northern fulmars were from 1975, 1987, 1993, and 2003. Between 8 and 10 individuals were analyzed per year. Analytes included C7-C15 perfluorinated carboxylates (PFCAs) and their suspected precursors, the 8:2 & 10:2 fluorotelomer saturated and unsaturated carboxylates (FTCAs, FTUCAs), C6, C8 (perfluorooctane sulfonate, PFOS), C10 sulfonates, and perfluorooctane sulfonamide (PFOSA). Liver samples were homogenized, liquid-liquid extracted with methyl tert-butyl ether, cleanedup using hexafluoropropanol, and analyzed by LC-MS/ MS. Overall, concentrations in seabirds were lower than those in other marine animals that occupy similar or higher trophic positions. In contrast to most other wildlife samples, PFC profiles were dominated by the PFCAs which comprised 81% and 93% of total PFC profiles in the 2004 thick-billed murre and 2003 northern fulmar samples, respectively. As well, the PFCA profiles were mainly comprised of the C11-C15 PFCAs, which appears to be unique among other wildlife species. PFC concentrations were found to increase significantly from 1975 to 2003/2004. Doubling times in thick-billed murres ranged from 2.3 yrs for perfluoropentadecanoate (PFPA) to 9.9 yrs for perfluorododecanoate (PFDoA), and from 2.5 yrs for PFPA to 11.7 yrs for perfluorodecanoate (PFDA) in northern fulmars. PFCA concentration increases in thick-billed murres were significant for both time periods (1975f1993, 1993f2004), but in northern fulmars appeared to remain steady after 1993. Differences in the temporal trends observed may be the result of differing migratory patterns of the seabirds. Finally, the detection of the 8:2 and 10:2 FTUCAs in seabirds is suggestive of fluorotelomer alcohols as a source of some PFCAs. * Corresponding author phone: (613) 998-6694, fax: (613) 9980458; e-mail:
[email protected]. † University of Toronto. ‡ Environment Canada. § National Wildlife Research Centre, Environment Canada. 10.1021/es062710w CCC: $37.00 Published on Web 04/18/2007
2007 American Chemical Society
Introduction The occurrence of perfluorinated alkyl compounds (PFCs) in wildlife and humans, including remote regions such as the Arctic, has been extensively reported (1-3). The most commonly monitored compounds comprise two PFC classes, the perfluorinated carboxylates (PFCAs) and the perfluorinated sulfonates (PFSAs). Although most studies have only analyzed for the eight-carbon perfluorinated acids, perfluorooctanoate (PFOA) and PFOS, there has been increasing attention concerning the longer-chain PFCAs. In fact, in most wildlife tissues the overall PFCA profile is dominated by the long-chain PFCAs (4) due to the greater bioaccumulation factors of these compounds (5, 6). There is some evidence of increased toxicological effects associated with increasing fluorinated chain-length, such as the inhibition of gapjunction intercellular communication (7). In addition, there have been several reports of known PFCA and PFSA precursors in wildlife tissues, such as the perfluoroalkyl sulfonamido alcohols, N-EtFOSA, and the FTCAs and FTUCAs (8-10). Fluorotelomer alcohols (FTOHs) have been shown to degrade to PFCAs, through FTCAs and FTUCAs, via atmospheric oxidation (11, 12) and biotic processes such as microbial degradation (13, 14) and rat metabolism (15). Perfluoroalkyl sulfonamido alcohols have been shown to degrade via atmospheric oxidation to PFCAs and PFSAs, in a mechanism analogous to FTOH oxidation (16, 17). Two mechanisms have been proposed to explain the transport of PFCs to the arctic environment: indirectly through atmospheric transport and directly via the ocean currents. The atmospheric mechanism involves the atmospheric transport of volatile precursor chemicals, such as FTOHs and polyfluorinated sulfonamido alcohols that degrade to PFCAs and PFSAs. FTOHs and polyfluorinated sulfonamido alcohols have been detected in the atmosphere (18-20). Further, it has been shown that the atmospheric lifetime of FTOHs with respect to hydroxyl radical reaction is ∼20 days, and is sufficiently long to permit transport to the Arctic (21). The other mechanism involves the direct transport of PFCAs themselves to the Arctic through oceanic transport (22, 23). PFCs are presumably transported to the Arctic through a combination of both mechanisms. However, the short PFOS disappearance half-life in ringed seals following 3M’s PFOSF phase-out (10) as well as the detection of almost exclusively linear isomers of several PFCAs (24) is consistent with the atmospheric transport of PFC precursors as a significant transport mechanism to the Arctic wildlife. Seabirds have previously been used for the biomonitoring of temporal trends in “legacy” (i.e,, PCBs, organochlorinated pesticides, mercury) and “new” (i.e., PBDEs) contaminants in the Arctic (25-27). The contaminant body burden in seabirds represents an integration of exposure from both the breeding grounds and overwintering locations. Thus, interpretation of temporal trends may be confounded by the migratory behavior of seabirds that overwinter in differing locations in addition to potentially differing diets and metabolic capacities. This may result in differing temporal trends of contaminants among seabird species (27). There have been relatively few reports of PFCs in seabirds from arctic regions. Previous studies include northern fulmars (4), black-legged kittiwakes (8), and glaucous gulls (8) from the Canadian Arctic, and glaucous gulls from the Norwegian Arctic (28). However, only two of the studies analyzed for the longer-chain PFCAs (4, 28). Not surprisingly, PFC concentrations in seabirds from these remote regions are lower than VOL. 41, NO. 10, 2007 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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those measured in species from industrial regions (3). PFC concentrations in fish-eating birds are generally lower than those in fish-eating mammals, which has been suggested to be due a shorter elimination half-life in birds (3). Temporal studies of PFCs in wildlife are limited, particularly those from arctic regions. In fact, we are not aware of any previous studies of PFC temporal trends in arctic seabirds. Overall, there appears to be a general increasing trend of PFCAs and PFOS concentrations in wildlife over time (3). However, a recent decline in PFOS levels in arctic ringed seals, from the early 2000s through 2005, has been observed (10). Regarding seabirds from non-arctic regions, white-tailed sea eagles from eastern Germany and Poland showed significant PFOS increases from 1979 to 1999 (29). Also, guillemot eggs from the Baltic Sea showed a 30-fold increase in PFOS concentrations between 1968 and 2000 (30). However, PFOS concentrations in these samples appeared to decline post-1997, although this trend was observed only in the pooled egg samples. This paper presents temporal trends of PFCs in two seabird species, thick-billed murres and northern fulmars, from Prince Leopold Island in the Canadian Arctic. The influence of overwintering location on temporal trends is described. Seabird liver samples were analyzed for C7-C15 PFCAs, the suspected PFCA precursors 8:2 and 10:2 saturated and unsaturated fluorotelomer acids, and C6, C8, and C10 PFSAs.
Materials and Methods Standards and Chemicals. Standards and reagents used were identical to those previously reported by our laboratory (10). Sample Collection. Adult northern fulmars (Fulmarus glacialis) and thick-billed murres (Uria lomvia) were collected from the Prince Leopold Island (PLI) Migratory Bird Sanctuary (74°02′ N, 90°05′ W) in Lancaster Sound, Nunavut, Canada (Figure S1 in the Supporting Information (SI)). Northern fulmars were collected in 1975 (n ) 9), 1987 (n ) 8), 1993 (n ) 10), and 2003 (n ) 9), whereas thick-billed murres were collected in 1975 (n ) 8), 1993 (n ) 10), and 2004 (n ) 10). All birds were collected using quick-kill techniques. With the exception of the 2003 fulmars, birds were shipped to the Canadian Wildlife Service (CWS) at the National Wildlife Research Centre (NWRC) in Gatineau, Quebec, where they were dissected and livers were removed under chemically clean conditions. Livers were homogenized, transferred to acetone-hexane rinsed glass vials, and stored at -40 °C. These storage conditions have been shown to avoid changes over time in the concentration of various organochlorine contaminants as well as moisture content (31, 32). Livers from the 2003 fulmars were excised in the field, transferred to a Whirlpak bag, placed in a second Whirlpak bag, and kept on ice until shipment back to the NWRC. Specific ages could not be determined, but all individuals were classified as adult birds. Sample Extraction and Instrumental Analysis. Liver samples were extracted using MTBE and hexafluoropropanol (HFP) as described in detail elsewhere (10, 33). Briefly, the MTBE extract was blown-down to ∼0.5 mL under a gentle stream of nitrogen gas, an equal volume of HFP was added, and the precipitated components were removed by filtering. The MTBE/HFP mixture was evaporated to dryness and reconstituted with 500 µL of methanol. The suite of internal standards was added immediately prior to instrumental analysis. Instrumental analysis was performed by liquid chromatography with negative electrospray tandem mass spectrometry (LC-MS/MS) under conditions previously described (33). Analytes were detected using an API 4000 Q Trap (Applied Biosystems/MDS Sciex, Concord, ON) with samples injected with an Agilent 1100 autosampler (injection volume ) 100 µL, flow rate ) 250 µL/min). Chromatography 3522
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was performed using a Genesis C18 column (50 × 2.1 mm, 4 µm particle size, Chromatographic Specialties, Brockville, ON), preceded by a C18 guard column (4.0 × 2.0 mm, Phenomenex, Torrance, CA). The mobile phase was methanol (MeOH)/water (0.01 M ammonium acetate) and analytes were separated using gradient conditions. Initial conditions were 75:25 MeOH/water, increasing to 90:10 MeOH/water over 3 min, followed by a 3 min hold, and then reverting to initial conditions. Target analytes included perfluoroheptanoate (PFHpA), PFOA, perfluorononanoate (PFNA), PFDA, perfluoroundecanoate (PFUnA), PFDoA, perfluorotridecanoate (PFTrA), perfluorotetradecanoate (PFTA), PFPA, 8:2 & 10:2 FTCA, 8:2 & 10:2 FTUCA, perfluorohexane sulfonate (PFHxS), PFOS, perfluorodecane sulfonate (PFDS), and PFOSA. Analyte responses were normalized to internal standard responses. The internal standard mix was added to give a final concentration of approximately 250 ng/L for all internal standards. 13C2-PFDA was used for all PFCAs, PFSAs, and PFOSA; 10:2 13C2-FTUCA was used for 10:2 FTCA and 10:2 FTUCA, and 8:2 13C2-FTUCA was used for 8:2 FTCA and 8:2 FTUCA. PFTrA was quantified using the response factor for PFDoA, and PFPA was quantified using the response factor for PFTA since analytical standards were not available. Concentrations were not corrected for recovery. Statistical Analysis and Data Treatment. Instrumental detection limits (IDL) were determined as the standard deviation from nine injections of the lowest calibration standard. Method detection limits (MDL) were determined as three times the standard deviation of the procedural blanks. The IDL was used if analytes were not detected in the blanks. Concentrations less than the IDL were reported as nondetect (nd), whereas concentrations less than the MDL were reported as
