Environmental Research 134 (2014) 46–56

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A geographical comparison of chlorinated, brominated and fluorinated compounds in seabirds breeding in the eastern Canadian Arctic$ Birgit M. Braune a,n, Anthony J. Gaston a, Robert J. Letcher a, H. Grant Gilchrist a, Mark L. Mallory b, Jennifer F. Provencher c a

Environment Canada, National Wildlife Research Centre, Carleton University, Raven Road, Ottawa, Ontario, Canada K1A 0H3 Biology Department, Acadia University, Wolfville, Nova Scotia, Canada B4P 2R6 c Department of Biology, Carleton University, Ottawa, Ontario, Canada K1S 5B6 b

art ic l e i nf o

a b s t r a c t

Article history: Received 15 May 2014 Received in revised form 19 June 2014 Accepted 20 June 2014

A suite of chlorinated, brominated and fluorinated organic contaminants were measured in livers of adult thick-billed murres (Uria lomvia) and northern fulmars (Fulmarus glacialis) from several locations in the eastern Canadian Arctic during 2007–2008. Thick-billed murres were collected from five colonies (Coats Island, Digges Island, Akpatok Island, Prince Leopold Island, Minarets) and northern fulmars from two colonies (Prince Leopold Island, Minarets). Legacy organochlorines (e.g. PCBs, DDT, chlorobenzenes, chlordanes) and perfluorooctane sulfonate (PFOS) dominated the compositional profiles of the measured halogenated compounds in the livers of both species at all colonies. Among the murre colonies sampled, Prince Leopold Island birds generally had the highest mean concentrations of organochlorines, whereas the highest mean concentration of sum (Σ) polybrominated diphenyl ethers (PBDEs) was found at the Minarets and the lowest at Prince Leopold Island. PBDEs were detected in only a few fulmar livers from the Minarets and in none of the fulmar livers from Prince Leopold Island. Mean PFOS concentrations were highest in both murre and fulmar livers at Prince Leopold Island. PFOS was approximately two orders of magnitude higher than the mean sum (Σ) perfluorinated carboxylate (PFCA) concentration in both species and at all colonies. The reasons for inter-colony and inter-species differences in contaminant liver levels are probably variable and complex, and likely reflect differences in contaminant transport and exposure pathways, as well as differences among colonies in their diets and overwintering areas. To our knowledge, this is the first spatial assessment of PBDEs, PFCAs and PFOS in seabirds from the Canadian Arctic. Crown Copyright & 2014 Published by Elsevier Inc. All rights reserved.

Keywords: Canadian Arctic Seabirds Organochlorines Polybrominated diphenyl ethers Perfluorinated compounds Spatial analysis

1. Introduction A growing number of persistent chlorinated, brominated and fluorinated organic contaminants have been shown to be ubiquitous in arctic biota including avian wildlife (Braune et al., 2005; Butt et al., 2010; de Wit et al., 2010; Letcher et al., 2010). These chemical contaminants are, for the most part, transported there by

☆ Funding sources: Financial support was provided by International Polar Year 2007–2009, Aboriginal Affairs and Northern Development Canada (Northern Contaminants Program and Northern Scientific Training Program), the Arctic Institute of North America (Grant-in-Aid Program), Natural Sciences and Engineering Research Council of Canada (NSERC), and the Nasivvik Centre for Inuit Health and Changing Environments which is supported by Université Laval and Trent University. Research approval: collections were made in accordance with guidelines from the Canadian Council on Animal Care, and under appropriate territorial and federal research permits. n Corresponding author. Fax: þ1 613 998 0458. E-mail address: [email protected] (B.M. Braune).

http://dx.doi.org/10.1016/j.envres.2014.06.019 0013-9351/Crown Copyright & 2014 Published by Elsevier Inc. All rights reserved.

air and/or ocean currents (Macdonald et al., 2000). Air is the most important transport route to the Arctic for volatile and semivolatile contaminants (Wania, 2003), but not for non-volatile compounds that include per- and poly-fluoroalkyl substances (PFASs) such as the perfluorinated sulfonates (PFSAs) and carboxylates (PFCAs). PFSAs and PFCAs of varying carbon chain lengths have been found in a wide variety of arctic wildlife (Butt et al., 2010; Houde et al., 2011; Letcher et al., 2010). It has been proposed that neutral, volatile precursor compounds of PFCAs and PFSAs, such as fluorotelomer alcohols (FTOHs) and sulfonamide alcohols, undergo long-range atmospheric transport and are degraded in remote regions (Ellis et al., 2004; Martin et al., 2006; Schenker et al., 2008; Young and Mabury, 2010). Alternatively, ionizable PFASs, such as the more water-soluble and less volatile PFSAs and PFCAs, could be transported directly to the arctic marine environment via ocean currents (Armitage et al., 2006; Wania, 2007). Upon reaching the Arctic, most persistent pollutants biomagnify through food webs making those species feeding at higher trophic levels more vulnerable to exposure via their diet (Borgå et

B.M. Braune et al. / Environmental Research 134 (2014) 46–56

al., 2004; Braune et al., 2005; Butt et al., 2010; de Wit et al., 2010; Hop et al., 2002; Houde et al., 2011; Vorkamp et al., 2004; Wolkers et al., 2004). Seabirds such as thick-billed murres (Uria lomvia) and northern fulmars (Fulmarus glacialis) feed at relatively high trophic positions in arctic marine food webs (Hobson et al., 2002; Hop et al., 2002) making them ideal sentinel species for a spatial analysis of persistent organohalogens. A spatial survey of contaminants in Canadian Arctic seabirds was carried out in 1993 and included thick-billed murres from four locations: two in the Canadian high Arctic, and two in northern Hudson Bay. Concentrations of many of the legacy persistent organic pollutants (POPs), such as PCBs and DDT, differed significantly in eggs of murres between one of the high Arctic colonies and the other colonies sampled (Braune et al., 2002). To our knowledge, no spatial data have been reported for PFASs or brominated compounds in Canadian Arctic seabirds. As part of a recent study assessing changes in diets of arctic marine birds (Mallory et al., 2010; Provencher et al., 2012), adult thick-billed murres and northern fulmars were collected from several locations in the eastern Canadian Arctic during 2007– 2008. In this paper, we present data on contaminants in livers from those birds to: (i) determine recent spatial patterns for a range of known and persistent organochlorines, and (ii) examine spatial patterns for suites of more recently discovered contaminants such as polybrominated diphenyl ethers (PBDEs), hexabromocyclododecane (HBCDD) and PFASs in seabirds.

2. Materials and methods 2.1. Sample collection During 2007–2008, adult thick-billed murres were collected from waters adjacent to five colonies in the eastern Canadian Arctic: Coats Island (621980 N, 821000 W), Digges Island (621330 N, 771350 W), Akpatok Island (601580 N, 681080 W), Prince Leopold Island (741020 N, 901000 W), and Akpait (also known as the “Minarets”) (671000 N, 611800 W) on eastern Baffin Island (Fig. 1). Adult northern fulmars were collected from two areas: Prince Leopold Island and waters adjacent to the breeding colonies at the Minarets and Cape Searle (671150 N, 621350 W), about 30 km apart. Sampling details can be found in Provencher et al. (2009, 2012) and Mallory et al. (2010). The murres from Coats Island and the Minarets were collected in late July–early August 2007, and the rest of the birds were collected in August 2008. All female birds were assumed to be breeding birds based on the presence of brood patches. Livers of five male and five female birds from each species and sampling location (except Coats Island) were removed and stored in acetone-hexane rinsed glass vials for subsequent organohalogen analysis. At Coats Island, only female murres were sampled in adequate numbers (n¼ 5 females). Liver samples were shipped to the National Wildlife Research Centre (NWRC) in Ottawa, Ontario, where they were homogenized and stored at  40 1C. All birds were taken under appropriate research and collection permits. 2.2. Chemical analyses 2.2.1. Organochlorines Liver homogenates were individually analyzed for organochlorines (OCs) including chlorobenzenes (ΣCBz¼ 1,2,4,5-tetrachlorobenzene, 1,2,3,4-tetrachlorobenzene, pentachlorobenzene and hexachlorobenzene), hexachlorocyclohexanes (ΣHCH ¼ α-, β- and γ-hexachlorocyclohexane), chlordane-related compounds (ΣCHL¼ oxychlordane, trans-chlordane, cis-chlordane, trans-nonachlor, cis-nonachlor and heptachlor epoxide), DDT and its metabolites (ΣDDT ¼p,p0 -DDE, p,p0 -DDD and p,p0 -DDT), octachlorostyrene (OCS), mirex, dieldrin and PCB congeners (ΣPCB). Measurements for 71 PCB congeners, as identified according to IUPAC numbers (Ballschmiter et al., 1992), were reported: 16/32, 17, 18, 22, 28, 31, 33/20, 42, 44, 47/ 48, 49, 52, 56/60, 64/41, 66, 70/76, 74, 85, 87, 92, 95, 97, 99, 101/90, 105, 110, 114, 118, 128, 130, 137, 138, 141, 146, 149, 151, 153, 156, 157, 158, 167, 170/190, 171, 172, 174, 176, 177, 178, 179, 180, 183, 187, 189, 194, 195, 196/203, 199, 200, 202, 206, 207, and 208. However, only 59 congeners were detected: 28, 31, 44, 47/48, 49, 52, 56/60, 64/41, 66, 70/76, 74, 85, 87, 92, 95, 99, 101/90, 105, 110, 114, 118, 128, 130, 137, 138, 141, 146, 149, 151, 153, 156, 157, 158, 167, 170/190, 171, 172, 177, 178, 179, 180, 183, 187, 189, 194, 195, 196/203, 199, 200, 206, 207, and 208. Congeners that are separated above by a slash chromatographically co-eluted during the separation process and are therefore reported together.

47

Samples were analyzed for organochlorines by gas chromatography using a mass selective detector (GC/MSD) and lipids were determined by gravimetric methods. Chemical extraction and cleanup of PCBs and organochlorine pesticides followed the procedures of Lazar et al. (1992). Briefly, tissue homogenates were ground with anhydrous sodium sulfate, spiked with labeled 13C-OC/PCB quantification standards and extracted with dichloromethane:hexane (50:50% v/v). Sample clean-up was performed by gel permeation chromatography followed by activated Florisil chromatography. Chemical analysis was performed using a capillary gas chromatograph (Agilent 6890N) coupled with a mass selective detector (Agilent 5973N) operated in electron impact (EI) mode. The column was a 30 m  0.25 mm  0.25 mm DB-5 column. PCBs and organochlorine pesticides were determined using an external quantification approach as described by Drouillard and Norstrom (2003). Three duplicate extractions, one duplicate injection, two method blanks and four in-house reference materials (DCCO Reference Egg Pools DCCOQA-2008-12 and DCCOQA2009-01, HERG Reference Egg Pools HERGQA-2008-20 and HERGQA-2009-01) were run for quality control. Internal standard recoveries averaged 9371.4% (mean7SE). Therefore, residues were not corrected for internal standard recoveries. The nominal detection limit was 0.1 ng g  1 wet weight (ww).

2.2.2. Brominated contaminants Liver homogenates were analyzed for PBDEs, two polybrominated biphenyls (PBBs) and total-(α)-HBCDD. Sample extraction and clean-up were the same as for the organochlorines except that tissue homogenates were spiked with an internal standard (BDE-30). Using this method, it has been shown the BDE-30 is a representative internal standard for the PBDEs, PBBs and HBCDD (Chen et al., 2012). Chemical analysis for 13 BDE congeners (BDE-17, -28, -47, -49, -66, -85, -99, -100, -138, -153, -154 (co-elution with BB-153), -183 and -190), BB-101 and total-(α)-HBCDD was performed using an Agilent 6890 gas chromatograph (GC) equipped with a 5973 quadrupole mass spectrometer (MS) detector run in electron capture negative ionization (ECNI) mode. The GC column was a 15 m  0.25 mm  0.1 mm DB-5HT capillary column. The determination of α-HBCDD by GC–MS is representative of total-HBCDD, as low levels of β- and γ-HBCDD isomer residues are thermally isomerized to α-HBCDD in the injection port at temperatures exceeding 160 1C (Chen et al., 2012). PBDEs, HBCDD and PBBs were identified on the basis of their retention times on the DB-5HT GC columns relative to authentic standards. Three duplicate extractions, one duplicate injection, two method blanks and the same four in-house reference materials identified for the organochlorine analysis were run for quality control. Quantification of the brominated compounds was performed using an internal standard method based on the relative ECNI response factor of the 79 Brþ 81Br anions of BDE-30 and to that of authentic congener standards in the neutral fractions. The method limit of quantification (MLOQ) was 0.1 ng g  1 ww for all PBDE and PBB congeners, and for HBCDD, the MLOQ was 1 ng g  1 ww. The recovery efficiency of BDE-30 averaged 1047 2.1% (mean 7SE). All reported residue levels were inherently corrected for recovery as the internal standard quantification approach was used. BDE-153 and BDE-154 have very similar physical–chemical characteristics (Tittlemier et al., 2002) and published data for biota show that BDE-154 generally occurs at lower or similar concentrations to BDE-153 (e.g., Elliott et al., 2005; Hites, 2004; McKinney et al., 2011; Norstrom et al., 2002). Given that BDE-153 was detected in only 11% of our samples compared with 60% for BDE-154/BB-153, and where both were quantified, concentrations of BDE-153 were consistently lower than for BDE-154/BB-153, we concluded that BDE-154/BB-153 was likely comprised of over 90% BB-153 and, therefore, was not included in the calculation of ΣPBDE concentrations. ΣPBDE was standardized to the sum of BDE-17, -28, -49, -47, -66, -100, -99, -85, -153, -138, -183 and -190.

2.2.3. Per- and poly-fluoroalkyl substances (PFASs) The PFAS extraction, cleanup and analysis have been described elsewhere (Gebbink and Letcher, 2012; Greaves et al., 2012). Briefly, approximately 1 g of liver homogenate was spiked with labeled internal standards and extracted with 10 mM KOH acetonitrile/water. The cleanup and fractionation of the overall PFAS extract was performed using Waters Oasis WAX solid phase extraction (SPE) cartridges. The first fraction contained FTOHs and FOSAs; the second fraction contained PFSAs, PFCAs and fluorotelomer unsaturated acids (FTUCAs). The separation of the target compounds in both fractions was carried out on a Waters 2695 HPLC equipped with an ACE 3 C18 analytical column (50 mm  2.1 mm I.D., 3 μm particle size, Advance Chromatography Technologies, Aberdeen, UK) coupled to a Waters Quattro Ultima triple quadrupole mass spectrometer (Waters, Milford, MA, USA). Analysis of PFCAs, PFSAs and FTUCAs was done using negative electrospray ionization (ESI  ), and the FTOHs and FOSAs were analyzed by negative atmospheric pressure photoionization (APPI  ). Quantification was performed using an internal standard approach. Standards for the PFSAs [C4 (PFBS), C6 (PFHxS), C8 (PFOS), C10 (PFDS)], PFCAs (C6–C15 chain lengths: PFHxA, PFHpA, PFOA, PFNA, PFDA, PFUnA, PFDoA, PFTrA, PFTeA and PFPA, respectively), 6:2, 8:2 and 10:2 FTUCAs, 6:2, 8:2 and 10:2 FTOHs, and two FOSAs [perfluorooctanesulfonamide (PFOSA), methylated perfluorooctanesulfonamide (N-Me-FOSA)] as well as all internal labeled standards were obtained from Wellington Laboratories (Guelph, ON, Canada). See Table S1 for a complete listing of all of the above PFASs as well as the 13C- or 18Oenriched internal standards used. All solvents used were HPLC grade and purchased

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B.M. Braune et al. / Environmental Research 134 (2014) 46–56

Fig. 1. Sampling locations and known overwintering areas of thick-billed murres (TBMU) and northern fulmars (NOFU) in the Canadian Arctic. Information sources for overwintering areas: Mallory et al. (2008) and McFarlane Tranquilla et al. (2013). from Fisher Scientific (Ottawa, Canada). The calibration curve for PFTeA was used for PFPA quantification since a PFPA standard was unavailable at the time of this study. Where no labeled standards were available, labeled internal standards with the closest retention time were used. Since an isotope dilution quantification approach was used, concentrations were inherently recovery-corrected. Recoveries of the FTOH and FOSA internal standards averaged 58% and, for the PFCAs, PFSAs and FTUCAs, recoveries averaged 56%. For every block of 10 samples, a blank sample and a NWRC in-house reference material (spiked pork liver) was analyzed. See Table S1 for limits of detection (LODs) and method quantification limits (MQLs) for the measured PFASs.

2.3. Data treatment To address the possibility of inter-year variation in contaminant concentrations among murres sampled over two years (i.e. murres sampled from Coats Island and the Minarets in 2007 and from the other three colonies in 2008), we used t-tests to compare concentrations of ΣCBz, OCS, ΣCHL, ΣDDT, dieldrin, ΣPCB, ΣPBDE, Hg and δ15N in eggs of thick-billed murres sampled from Coats Island in 2007 (n¼ 15) and 2008 (n ¼15) (Braune, unpublished data). There were no significant inter-year differences found for the contaminant groups analyzed or for δ15N or % lipid in the murre eggs sampled from Coats Island. This suggests that there was no significant difference in contaminant exposure of the murres or in murre diet between the two years. Unfortunately, similar data were not available for the Minarets from those two years. However, based on the results for Coats Island, we have assumed that the contaminants data for the murres sampled from the two years are comparable. In order to determine which of the major halogenated groups/compounds dominated the measured contaminant profile, wet-weight concentrations were used to calculate the percent contribution of the major halogenated groups/ compounds to the sum total measured (i.e. sum of ΣOCþ PFOSþ ΣPFCA þ ΣPBDE). Total organochlorine (ΣOC) concentrations were calculated as the sum of ΣCBz þ ΣHCHþ ΣCHL þ ΣDDTþ OCSþ mirex þdieldrin þ Σ59PCB. ΣPFCA was the sum of the C6–C15 chain lengths. All statistical tests were performed using Statistica for Windows Version 7.0 (StatSoft Inc., Tulsa, OK) with a significance level of po 0.05. Only those analytes for which 470% of the samples had quantifiable concentrations for a given species

were statistically analyzed. Non-detect values were set to one half the detection limit for purposes of statistical analyses but were set to zero for calculation of the sums of major halogenated groups (e.g. Σ59PCB, ΣDDT, ΣPBDE, ΣPFCA). Tests for normality and sample variance in the data were performed using the Shapiro– Wilks' W and Levene's tests, respectively. Data which violated the assumptions of normality and homogeneity of sample variance were loge-transformed, as required (Zar, 1984). Mann–Whitney U-tests were used to compare compositional profiles (by percent) of ΣOC as well as the major organochlorine groups (e.g. Σ59PCB, ΣDDT, ΣCBz, ΣCHL, ΣHCH) between males and females of thick-billed murres at the four colonies where both sexes were sampled, and for northern fulmars from the two colonies sampled as well as between fulmar colonies. Kruskal–Wallis analysis of variance (ANOVA) was used to test for differences in compositional profiles (by percent) among murre colonies. Northern fulmars are sexually dimorphic and since tarsal length is the best single-variable character in Arctic fulmar populations (Mallory and Forbes, 2005), a body condition index was calculated individually for fulmars based on the ratio of body mass to tarsal length (Table S2). Relationships of the body condition index with colony and sex were tested using factorial analysis of variance (ANOVA) followed by Tukey's pairwise multiple comparison test. Results showed no significant difference between sexes or colonies for the body condition index, nor did % lipid in liver vary significantly by colony or sex. Regression analyses indicated no significant relationship of body condition with % lipid in liver in either males or females analyzed separately or combined, nor did % lipid vary by colony or sex. However, there was a significant relationship between % lipid and concentration for most organochlorine contaminants. As well, body condition index varied significantly with concentrations of a few contaminants in female fulmars. Therefore, further statistical analyses of contaminant concentrations in fulmars were carried out using lipid-normalized data and included the body condition index as a covariate. In thick-billed murres, sexual dimorphism is slight and previous studies have shown no systematic difference in body mass between the sexes (Gaston and Nettleship, 1981; Gaston and Hipfner, 2000). A factorial ANOVA testing for effects of colony or sex relative to body mass for the four colonies for which we had sampled both males and females showed no significant effect of sex on body mass but did indicate a significant difference (p ¼ 0.003) among colonies. Percent lipid in liver also did not vary significantly by colony or sex. Regression analyses indicated no

B.M. Braune et al. / Environmental Research 134 (2014) 46–56 significant relationship of body mass with % lipid in liver in either males or females analyzed separately or combined. However, there was a significant relationship between % lipid and concentrations of most contaminants, as well as a significant relationship between body mass and concentrations of some contaminants. Given the relatively small amount of body mass variation explained by linear measurements in previous studies (Gaston and Hipfner, 2006; Jacobs et al., 2012), we used raw body mass (Table S3) as our covariate in subsequent statistical analyses of the lipid-normalized contaminant concentration data. Statistical analyses comparing residue concentrations between sexes, among colonies, and between species for all quantifiable organochlorines and PBDEs were carried out using lipid-normalized data. However, since PFCAs and PFSAs are generally associated with proteins rather than lipid (Butt et al., 2010; Greaves et al., 2012), those concentrations were not lipid-normalized. Differences in contaminant concentrations between sexes and between colonies for the fulmars were tested using analyses of covariance (ANCOVA) controlling for body condition index followed by Tukey's post hoc pairwise multiple comparison tests. There were no significant effects of sex or the sex  colony interaction term for any of the contaminant concentrations tested except for ΣHCH. There was a significant colony  sex effect for ΣHCH (p ¼ 0.03) but Tukey's post hoc test showed no differences between sexes by colony. Therefore, inter-colony differences in contaminant concentrations were tested with sexes pooled using ANCOVA. For the murres, ANCOVA controlling for body mass was used to test for differences in contaminant concentrations between sexes and among colonies for all colonies except Coats Island. Since there were no significant effects of sex or the colony  sex interaction term for any of the contaminant concentrations tested, inter-colony differences in contaminant concentrations for all five murre colonies were tested with sexes pooled using ANCOVA followed by Tukey's post hoc pairwise multiple comparison tests for unequal sample sizes. Differences in contaminant concentrations between species were tested using factorial ANOVA followed by Tukey's post hoc pairwise multiple comparison tests for effects of species, sex, colony, and the various interaction terms for the two colonies in our study where the two species co-occur (i.e. Minarets, Prince Leopold Island). The tabulated data are presented as arithmetic means7 one standard error (SE) in concentration units of ng g  1 lipid weight (lw) for the OCs and PBDEs, and ng g  1 ww for the PFASs.

3. Results There were no significant differences in hepatic residue concentrations of organochlorines (HCB, ΣCBz, β-HCH, ΣHCH, OCS, heptachlor epoxide, oxychlordane, cis-nonachlor, ΣCHL, p,p0 -DDE, ΣDDT, mirex, ΣPCB), ΣPBDE, BDE-47, PFOS and ΣPFCA between sexes of thick-billed murres at any of the four murre colonies where both males and females were sampled. There were also no significant differences in hepatic concentrations of organochlorines (HCB, ΣCBz, β-HCH, ΣHCH, OCS, heptachlor epoxide, oxychlordane, trans-nonachlor, ΣCHL, p,p0 -DDE, ΣDDT, mirex, ΣPCB), PFOS and ΣPFCA between sexes of northern fulmars at the two colonies sampled. The compositional profiles of ΣOC or profiles of any of the major organochlorine groups (e.g. Σ59PCB, ΣDDT, ΣCBz, ΣCHL, ΣHCH) also did not differ significantly between the sexes at any of the four murre colonies, nor for northern fulmars at the two colonies sampled. Therefore, data for males and females were combined. Of the major halogenated groups/compounds measured, ΣOC and PFOS dominated the compositional profiles of both species at all colonies with ΣPBDE contributing o 1% (Fig. 2A). Hepatic residue concentrations by species and colony are presented in Tables 1, S4 and S5. 3.1. Organochlorines The ΣOC profiles were dominated by Σ59PCB and ΣDDT followed by ΣCBz in thick-billed murres at all colonies whereas in the fulmars, Σ59PCB followed by ΣCHL and ΣDDT dominated the profiles (Fig. 2B). ΣDDT was composed primarily of p,p0 -DDE (colony averages: 97–100%) and HCB was the major chlorobenzene found (colony averages: 88–93%) in both murres and fulmars. The ΣCHL profile varied significantly among the murre colonies with oxychlordane dominating at Prince Leopold, Digges and Akpatok

49

Islands, and heptachlor epoxide dominating at Coats Island and the Minarets (Fig. 2C). Oxychlordane dominated the ΣCHL profile in the fulmars at both Prince Leopold Island and the Minarets (Fig. 2C). The Σ59PCB profile was dominated by the hexa-CB homolog followed by the penta- and hepta-homologs (Fig. S1). CB-153, -138 and -118 comprised over 35% of Σ59PCB in the murres and CB-153, -138 and -180 comprised over 50% of Σ59PCB in the fulmars. β-HCH was the major HCH isomer in both the murres (colony averages: 61–90%) and the fulmars (93%, 100%). 3.1.1. Inter-colony variation Organochlorine concentrations did not vary significantly between the two fulmar colonies. Only concentrations of mirex, p,p0 -DDE, ΣDDT and Σ59PCB did not vary significantly among murre colonies. For those organochlorines which did vary significantly among murre colonies, the murres from Prince Leopold Island had the highest concentrations (Table 2). At the two colonies (Prince Leopold Island, Minarets) where both fulmars and murres co-occur, there were no inter-species differences found for HCB, ΣCBz or dieldrin, whereas fulmars had significantly higher concentrations of OCS, chlordane-related compounds (oxychlordane, HE, ΣCHL), mirex, p,p0 -DDE, ΣDDT and Σ59PCB at both colonies (Table 3 and Fig. S2). β-HCH and ΣHCH were the only compounds which were significantly higher in the murres and only at Prince Leopold Island (Tables 1 and 3). 3.2. Brominated contaminants PBDEs were not detected in any fulmar livers from Prince Leopold Island and in only three fulmars from the Minarets (Table 1), but they were detected in 80% of the murres analyzed across all colonies. Total-(α)-HBCDD was detected only at the Minarets in four male murres (8.10–159 ng g  1 lw) and three male fulmars (8.30–49.3 ng g  1 lw), and BB-101 was detected in only six murres from the Minarets (2.29–28.7 ng g  1 lw) (Table S4). ΣPBDE varied significantly among the murre colonies with the highest concentrations found at the Minarets and lowest at Prince Leopold Island (Table 2). BDE-47 was the predominant BDE congener in the murres followed by BDE-100 and BDE-99 (Fig. 2D). At the Minarets, the PBDE profile in murres consisted of a larger proportion of BDE-100 as well as other BDE congeners, primarily BDE-138 and BDE-153. The PBDE profile in fulmars at the Minarets was dominated by BDE-153. 3.3. Per- and poly-fluoroalkyl substances PFOS was detected in all birds analyzed and was the major PFSA measured in both murres and fulmars (Table 1). PFOS averaged over 95% of ΣPFSA in the murres and over 99% ΣPFSA in the fulmars (Table S5). PFOSA was detected in only one fulmar from the Minarets (1.6 ng  1 g ww) and N-Me-FOSA was not detected in any of the samples (o0.2 ng g  1 ww). No FTOHs (6:2 FTOH, 8:2 FTOH, 10:2 FTOH) were detected ( o0.6 ng g  1 ww, o0.6 ng g  1 ww and o0.5 ng g  1 ww, respectively) in any of the samples, and FTUCAs (6:2 FTUCA, 8:2 FTUCA,10:2 FTUCA) were detected in a few samples but could not be quantified (o0.1 ng g  1 ww). PFOS was approximately two orders of magnitude higher than ΣPFCA in both species and at all colonies (Table 1). PFOS did not vary significantly among the murre colonies but fulmars from Prince Leopold Island had significantly higher levels of PFOS (p ¼0.02) than fulmars from the Minarets. ΣPFCA varied significantly among the murre colonies with the highest concentrations found at Coats Island (Table 2). There were no significant differences in ΣPFCA concentrations in fulmars between Prince Leopold

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B.M. Braune et al. / Environmental Research 134 (2014) 46–56

Thick-billed Murres

Digges

58%

41%

Akpatok

29%

71%

Coats

35%

60%

Minarets

53%

45%

ΣPFCA

Prince Leopold

44%

55%

ΣPBDE

Minarets

61%

38%

Prince Leopold

29%

70%

ΣOC PFOS

Northern Fulmars

Thick-billed Murres

Digges

37%

39%

Akpatok

38%

40%

Coats

37%

31%

Minarets

39%

39%

Prince Leopold

26%

32%

13% 13%

ΣPCB

25%

ΣDDT ΣCBz

17%

ΣCHL

25%

ΣHCH Northern Fulmars

Minarets

42%

24%

27%

Prince Leopold

35%

22%

31%

Dieldrin Mirex OCS

Thick-billed Murres

Digges

25%

69%

Akpatok

37%

57%

Coats

56%

13%

26%

Minarets

57%

20%

18%

Prince Leopold

27%

65%

heptachlor epoxide oxychlordane trans-nonachlor cis-nonachlor cis+trans-chlordane

Northern Fulmars

Minarets

8%

89%

Prince Leopold

8%

91%

Murres

Digges

77%

Akpatok

84%

Coats

87%

Minarets

54%

Prince Leopold

93%

BDE-47 BDE-99 BDE-100 BDE-other

13% 17%

18%

Fulmars

Minarets

11%

0

84%

20

40 60 80 Percent Composition

100

Fig. 2. Percent contributions of (A) ΣOC, PFOS, ΣPFCA and ΣPBDE to the total sum of halogenated compounds/groups (i.e. sum of ΣOC þPFOS þΣPFCA þΣPBDE), (B) PCBs (ΣPCB), DDT metabolites (ΣDDT), chlorobenzenes (ΣCBz), chlordane-related compounds (ΣCHL), hexachlorocyclohexanes (ΣHCH) dieldrin, mirex and octachlorostyrene (OCS) to the total organochlorines, (C) chlordane-related compounds to ΣCHL, and (D) of BDE congeners to ΣPBDE measured in livers of thick-billed murres and northern fulmars from the Canadian Arctic.

B.M. Braune et al. / Environmental Research 134 (2014) 46–56

51

Table 1 Mean ( 7 SE) concentrations of organochlorines (ng g  1 lipid weight), polybrominated diphenyl ethers (ng g  1 lipid weight) and perfluorinated compounds (ng g  1 wet weight) in livers of adult thick-billed murres from five colonies and northern fulmars from two colonies in the Canadian Arctic, 2007–08. Thick-billed Murres

% Lipid HCB ΣCBz β-HCH ΣHCH OCS Oxychlordane cis-Nonachlor trans-Nonachlor Heptachlor epoxide ΣCHL p,p0 -DDE ΣDDT Mirex Dieldrin Σ59PCB ΣPBDE PFOS ΣPFCA

Northern Fulmars

Coats Island n¼ 5

Digges Island n¼ 10

Akpatok Island n ¼10

Minarets n¼ 10

Prince Leopold Island n¼ 10

Minarets n¼ 10

Prince Leopold Island n¼ 10

1.2 70.4 893 7110 964 7122 38.2 723.4 50.9 731.2 30.3 75.74 36.0 736.0 28.8 75.01 2.27 72.27 65.5 78.37 137 738.2 1173 7163 1175 7161 21.1 711.0 19.8 719.8 1454 7280 8.05 73.45 104 730.4 5.78 70.76

2.8 7 0.4 555 7 106 6167 115 98.0 7 18.3 1397 21.6 26.5 7 4.68 1767 29.4 15.6 7 3.21 0.22 7 0.22 62.6 7 10.7 255 7 42.5 18357 334 18357 334 28.6 7 4.52 39.4 7 30.1 16407 255 19.3 7 3.97 1137 30.6 1.127 0.28

2.4 7 0.3 3587 58.3 3977 63.1 1017 19.1 1117 19.0 13.4 7 3.41 97.5 7 23.6 10.17 2.42 nd 56.3 7 13.1 1647 37.2 12727 221 12727 221 22.8 7 7.16 14.7 7 14.7 13447 325 11.3 7 4.68 230 7 40.9 0.60 7 0.19

2.17 0.5 1063 7 351 1202 7 391 33.67 18.9 48.77 25.8 22.0 7 5.31 44.0 7 26.5 35.57 10.6 6.62 7 2.60 1167 39.0 204 7 53.4 3053 7 952 31257 1002 12.4 7 6.91 16.3 7 12.2 28977 945 75.17 21.4 1207 36.4 2.58 7 0.33

2.3 7 0.6 1645 7 324 1858 7 351 2737 63.5 288 7 61.6 39.0 7 8.71 360 7 67.7 38.5 7 7.95 0.277 0.27 1517 31.7 550 7 105 24167 493 24167 493 25.7 7 7.27 206 7 52.3 1842 7 373 4.46 7 1.96 289 7 87.9 1.99 7 0.19

3.17 1.4 985 7 111 1084 7 120 83.4 7 10.1 90.3 7 11.0 96.0 7 20.0 6584 7 1379 2.047 2.04 81.5 7 20.1 553 7 90.4 72317 1418 5908 7 667 60487 674 2427 25.5 3447 129 11,1907 1446 5.46 7 2.93 349 7 105 4.94 7 1.89

1.0 7 0.1 14407 210 16017 233 99.7 7 14.3 99.7 7 14.3 1777 38.5 7660 7 1268 nd 50.9 7 9.62 6747 138 8388 7 1398 5563 7 984 5759 7 1026 2117 37.8 795 7 164 91847 1449 nd 6577 118 1.85 7 0.31

nd – not detected.

Table 2 Comparison of hepatic organohalogen concentrations in thick-billed murres among colonies. Different letters indicate statistically different residue concentrations among colonies for a given compound. The letter “A” indicates the colony (or colonies) with the highest concentration(s) and concentrations decrease significantly with each subsequent letter. The colony with the highest concentration is indicated by an asterisk and the colony with the lowest concentration is indicated by an underlined letter.

HCB ΣCBz β-HCH ΣHCH OCS Oxychlordane cis-Nonachlor Heptachlor epoxide ΣCHL ΣPBDE ΣPFCA

Coats Island n ¼5

Digges Island n¼10

Akpatok Island n ¼10

Prince Leopold Minarets Island n¼ 10 n¼ 10

ABC ABC AB B AB BC ABC

BC BC A B AB A BC

C C A B B AB C

AB AB B B AB C AB

An An An An An An An

AB

AB

B

AB

An

B AB An

AB AB CD

B B D

B An B

An B BC

Table 3 Inter-species comparison of hepatic organohalogen concentrations in thick-billed murres (TBMU) and northern fulmars (NOFU) from two colonies. Species with significantly higher concentrations indicated. Non-significant differences (p o 0.05) indicated by “ns”.

HCB ΣCBz Dieldrin β-HCH ΣHCH OCS Oxychlordane Heptachlor epoxide ΣCHL p,p0 -DDE ΣDDT Mirex Σ59PCB PFOS ΣPFCA

Minarets

Prince Leopold Island

ns ns ns ns ns NOFU 4TBMU NOFU 4TBMU NOFU 4TBMU NOFU 4TBMU NOFU 4TBMU NOFU 4TBMU NOFU 4TBMU NOFU 4TBMU ns ns

ns ns ns TBMU4 NOFU TBMU4 NOFU NOFU 4TBMU NOFU 4TBMU NOFU 4TBMU NOFU 4TBMU NOFU 4TBMU NOFU 4TBMU NOFU 4TBMU NOFU 4TBMU ns ns

4. Discussion 4.1. Organochlorines Island and the Minarets. However, although ANCOVA results indicated that PFOS concentrations were not significantly different between the two fulmar colonies (p ¼0.08), Tukey's post hoc test indicated that fulmars at Prince Leopold Island had significantly higher PFOS concentrations than those at the Minarets. There were no significant inter-species differences in PFOS and ΣPFCA concentrations in murres and fulmars at either Prince Leopold Island or the Minarets (Table 3). PFUnA (C11) was the dominant PFCA measured in livers of both murres and fulmars followed by either PFNA (C9) or PFTrA (C13) at most colonies (Table S5 and Fig. S3). Patterns varied among colonies and, at some colonies, between sexes, most notably for murres at Akpatok Island and for fulmars at Prince Leopold Island where PFDA (C10) was prominent in the females (Fig. S3).

Although there were some similarities in organochlorine concentrations across a number of the murre colonies sampled, adult murres from Prince Leopold Island had the highest hepatic concentrations of the organochlorines measured compared with the other murre colonies and, except for the HCHs, Akpatok and Coats Islands had the lowest (Table 2). Exposure to contaminants may vary between high and low Arctic areas due to different source regions, contaminant transport pathways, season and physical–chemical properties of contaminants (Bard, 1999; Gouin and Wania, 2007; Hallanger et al., 2011a, 2011b, 2011c; Ma et al., 2004; Macdonald et al., 2005; Shunthirasingham et al., 2010). Changes in food web structure may also affect an organism's exposure to contaminants (Macdonald et al., 2005). Historically, arctic cod (Boreogadus saida), which is a cold-water fish associated

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with ice cover, dominated the diet of high Arctic thick-billed murres while low Arctic murres also consumed arctic cod along with a number of secondary prey items (Gaston and Bradstreet, 1993). As a result of changing ice conditions, particularly in Hudson Bay – Hudson Strait, there has been a shift in murre diet since the 1970s and 1980s. Not only has the proportion of fish in the diet of low Arctic murres decreased, there has also been a shift from arctic cod to capelin (Mallotus villosus) and sandlance (Ammodytes sp.) (Gaston et al., 2003, 2012). In contrast, the proportion of fish in the murre diet has not changed at the Minarets and may have actually increased at Prince Leopold Island (Provencher et al., 2012), a dietary trend also noted for northern fulmars at Prince Leopold Island (Mallory et al., 2010). Northern fulmars sampled from the Minarets and Prince Leopold Island in 2008 showed no significant inter-colony differences in organochlorines measured which is consistent with the finding that fulmars sampled at those same two high Arctic locations in 2008 had similar diets (Mallory et al., 2010). There were, however, significant inter-species differences in hepatic concentrations of some organochlorines between murres and fulmars at the Minarets and Prince Leopold Island. In particular, fulmars had ΣCHL concentrations, primarily oxychlordane, which were an order of magnitude higher than in the murres, an observation consistent with results from the Northwater Polynya in northern Baffin Bay (Fisk et al., 2001). It has been suggested that procellariids, such as the northern fulmar, have a high metabolic capacity for biotransforming cis- and trans-chlordane into oxychlordane (Fisk et al., 2001), whereas murres are more efficient at metabolizing and eliminating chlordanes (Fisk et al., 2001; Kawano et al., 1988) which would explain the different patterns found for the two species in our study. What remains to be explained is why ΣCHL profiles varied among the murre colonies sampled (Fig. 2C). Apparently murres from Coats Island and the Minarets experienced greater exposure to heptachlor epoxide and to some extent, cis-nonachlor, than those at the other three colonies although the variation among colonies was small (Table 2). β-HCH was the major HCH isomer found in both murres and fulmars at all sampling locations in our study. HCH isomers are subject to biotransformation and seabirds appear to readily metabolize the γ- and α-isomers whereas the recalcitrant β-isomer biomagnifies in the food web (Borgå et al., 2004; Hop et al., 2002; Moisey et al., 2001). Concentrations of β-HCH were highest in the murres at Prince Leopold Island and lowest at the Minarets and Coats Island (Table 1) although only the murres from the Minarets differed significantly from the other four colonies (Table 2). β-HCH concentrations in fulmars were also higher at Prince Leopold Island than at the Minarets, although the difference was not statistically significant. The technical HCH insecticide formulation, primarily composed of α-, β- and γ-HCH isomers, is now virtually out of use worldwide but was heavily used in China, India and other Asian countries up until about 1990 (Wu et al., 2010). Due to differences in their physical–chemical characteristics, the α- and γ-HCH isomers have been shown to reach the Arctic mainly through atmospheric transport whereas β-HCH was transported mainly by ocean currents via the Bering Strait, thus delaying its arrival in the Arctic (Li and Macdonald, 2005). Consistent with the recalcitrant nature of β-HCH in biota and its pathway of entry into the Arctic region, β-HCH concentrations in biota decrease from the Bering-Chukchi Sea region eastward (Borgå et al., 2005; Braune, 2007; Li and Macdonald, 2005; Muir et al., 2000). Given the easterly flow of ocean currents through the Canadian Arctic Archipelago (Bidleman et al., 2007), β-HCH should reach Prince Leopold Island first among the murre colonies sampled (Fig. 3). The currents then flow from Foxe Basin into Hudson Strait with only a small flow entering Hudson Bay where it

mixes with Hudson Bay waters as they flow counter-clockwise around the bay (Bidleman et al., 2007). This suggests that the food webs near Digges Island and Akpatok Island, which lie in Hudson Strait, may experience greater exposure to β-HCH than those near Coats Island in northern Hudson Bay. As the water currents originating from the Pacific flow south along the east coast of Baffin Island (Fig. 3), they start to mix with Atlantic waters by the time they reach Davis Strait (Jones et al., 2003). Given that the waters of the western North Atlantic have lower HCH concentrations than Canadian Arctic waters (Bidleman et al., 2007), this would suggest that HCH concentrations are reduced as a result of mixing in Davis Strait near where the Minarets are situated. After breeding, the murres from the Minarets move rapidly to their wintering area (Fig. 1) between Newfoundland and southern Greenland (Gaston et al., 2011), an area which should also be subject to lower HCH exposure based on water circulation models. This could explain why, in the present study, murres from the Minarets contained low concentrations of β-HCH relative to Prince Leopold Island as well as Akpatok and Digges Islands. It does not explain why there was such little difference between the two fulmar colonies sampled, but it may well be that birds from these colonies winter in the same area in the North Atlantic (Mallory et al., 2008). As was the case for fulmars sampled in 2008, levels of the major legacy organochlorine groups in livers of adult northern fulmars sampled from Prince Leopold Island in July 2003 (postegg-laying) were also not statistically different between the sexes (Braune et al., 2010). Hepatic concentrations of the major legacy organochlorines in northern fulmars collected from Prince Leopold Island and the Minarets in 2008 averaged 3–5 times higher than levels found in fulmars from Prince Leopold Island and Cape Vera in 2003 (Braune et al., 2010), and more than an order of magnitude higher than those found in two northern fulmars sampled in 2001 from the Qikiqtarjuaq area near the Minarets (Mallory et al., 2005). As well, fulmar concentrations from our study were higher than levels found in fulmars sampled from the Northwater Polynya in Baffin Bay in 1998 (Buckman et al., 2004), in some cases (e.g. ΣCBz, ΣHCH, ΣCHL, ΣPCB) by two-fold or more. Likewise, most of the hepatic organochlorine concentrations measured in murres from high Arctic colonies in 2007–2008 were higher than levels in thick-billed murres sampled from the Northwater Polynya in Baffin Bay in 1998 (Buckman et al., 2004). In particular, ΣCBz, ΣDDT and ΣPCB were 2–4 times higher in murres sampled in 2007–2008. The reasons for the differences among these studies are not clear but may be partially related to diet given that δ15N values for the birds varied among the studies with mean values of 15.0‰ and 15.3‰ for fulmars from the Minarets and Prince Leopold Island, respectively, in 2008 (Braune et al., 2014) compared with 13.8‰ for fulmars from Prince Leopold Island in 2003 (Braune et al., 2010) and 14.0‰ in fulmars from the Northwater Polynya in 1998 (Buckman et al., 2004). Likewise, the murres from the Minarets and Prince Leopold Island had higher δ15N values (14.7‰ and 15.8‰, respectively) in 2008 (Braune et al., 2014) than murres from the Northwater Polynya (13.8‰) in 1998 (Buckman et al., 2004). This merits further examination. 4.2. Brominated contaminants To our knowledge, this is the first spatial examination of PBDEs in seabirds in the Canadian Arctic. ΣPBDE levels in fulmars from Prince Leopold Island were below the detection limit and the mean ΣPBDE concentration in fulmars from the Minarets (5.5 ng g  1 lw) was quite low, in fact, three orders of magnitude lower than hepatic ΣPBDE concentrations found in adult fulmars from Bjørnøya in the Barents Sea (5255 ng g  1 lw; Knudsen et al., 2007). Among the murres, mean hepatic ΣPBDE concentrations

B.M. Braune et al. / Environmental Research 134 (2014) 46–56

53

Fig. 3. Major surface currents and water flow through the Canadian Arctic. Information sources: Hogan (2009) and Prinsenberg (1985).

were lowest at Prince Leopold Island (4.5 ng g  1 lw) and an order of magnitude higher at the Minarets (75 ng g  1 lw) where ΣPBDE concentrations measured in four of the murres ranged as high as 121–204 ng g  1 lw. The mean ΣPBDE level at the Minarets was similar to that found in livers of thick-billed murres sampled from the Barents Sea in 2002 (66 ng g  1 lw; Savinova et al., 2007; t18 ¼0.83, p ¼0.42). These data suggest that murres from the Minarets experienced a higher exposure to PBDEs than birds from the other murre colonies. Unlike most thick-billed murres breeding in the Canadian Arctic which overwinter in the Davis Strait and northern Labrador Sea area (Fig. 1), thick-billed murres from the Minarets tend to overwinter farther south over the eastern Grand Banks and southern Labrador Sea (Gaston et al., 2011; McFarlane Tranquilla et al., 2013). These waters are influenced by the Gulf Stream and North Atlantic Drift which receive fluvial inputs from densely populated areas of the eastern North American seaboard. As well, the Grand Banks and southern Labrador Sea may be more affected than more northerly latitudes by the prevailing winds passing over the industrialized regions of eastern North America. These ocean and air pathways may transport PBDEs released into the environment via, for example, sewage sludge and landfills containing discarded flame retardant-containing products (Shaw and Kannan, 2009). Therefore, differences in overwintering areas may partially explain the higher ΣPBDE concentrations found in the murres from the Minarets. Fulmars from the Minarets, which had relatively low ΣPBDE levels, may overwinter anywhere across the North Atlantic (Hatch and Nettleship, 1998; Mallory et al., 2008), including areas potentially closer to European sources where PBDE use was much less than in North America (de Wit et al., 2010). The PBDE congener pattern of the murres from the Minarets (Fig. 2D) was very similar to that found for eggs of ivory gulls (Pagophila eburnea) collected from the Canadian high Arctic in 2004 (Braune et al., 2007). This pattern suggests exposure mainly to the commercial Penta-BDE mixture used in North America which was comprised primarily of BDE-47, -99 and -100 with

smaller contributions from BDE-153 (Hale et al., 2003). In contrast, BDE-153 predominated in the fulmar PBDE profile from the Minarets which may be related to a combination of its high biomagnification and low biotransformation potential (Chen et al., 2010) as well as its potential for long-range transport (Wania and Dugani, 2003). BDE-153 was also a major contributor to the PBDE profile found in fulmar muscle and eggs from the Faroe Islands in the Northeast Atlantic (Fängström et al., 2005). Although the use of the commercial Penta- and Octa-BDE mixtures has been banned or phased out in both Europe and North America (de Wit et al., 2010), the disposal of existing stores of PBDEs in the technosphere threatens to expose marine and other wildlife to PBDEs for years to come (Shaw and Kannan, 2009). 4.3. Per- and poly-fluoroalkyl substances This is the first spatial analysis, to our knowledge, of the major bioaccumulative PFASs, PFCAs and PFOS, in seabirds from the Canadian Arctic. Liver concentrations of PFOS did not vary significantly among the murre colonies, but for both murres and fulmars, PFOS levels were approximately two-fold higher at Prince Leopold Island than the Minarets, significantly so in the fulmars. In contrast, ΣPFCA varied significantly among the murre colonies although there were no significant differences in ΣPFCA concentrations in murres or fulmars between Prince Leopold Island and the Minarets. No significant inter-species differences were found for hepatic PFOS or ΣPFCA concentrations at either Prince Leopold Island or the Minarets although mean PFOS concentrations were at least two-fold higher in the fulmars at both colonies. In contrast, a comparison of PFOS and ΣPFCA in eggs of northern fulmars and thick-billed murres sampled from Prince Leopold Island in 2008 showed significantly higher ΣPFCA concentrations in the fulmar eggs with little difference in PFOS levels between the two species (Braune and Letcher, 2013). It is tempting to suggest that the disparity in the inter-species relationship of ΣPFCA in egg and liver may reflect the preferential accumulation of longer-chain length

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PFCAs in the egg yolk as demonstrated by Gebbink and Letcher (2012) for herring gulls (Larus argentatus). The PFCA profile in livers of female murres and fulmars from Prince Leopold Island (Fig. S3) was dominated by C9–C11 PFCAs, mainly PFUnA (C11) followed by PFDA (C10) in the fulmars and PFNA (C9) in the murres, whereas the PFCA profile found in the eggs (Braune and Letcher, 2013) was dominated by the longerchained PFCAs, mainly PFTrA (C13) followed by PFUnA (C11) in the fulmars and by PFUnA (C11) followed by PFTrA (C13) in the murres. Therefore, although the different PFCA profiles in livers and eggs of fulmars and murres are supported by the findings of Gebbink and Letcher (2012), the slight differences in hepatic PFCA profiles cannot explain the inter-species difference found in the eggs given the similarity in hepatic ΣPFCA concentrations found for murres and fulmars at Prince Leopold Island in our study. The PFCA profile in fulmar and murre livers sampled from Prince Leopold Island in 2003–2004 was dominated by the C11–C15 carboxylates with PFTrA (C13) predominating, followed by PFTeA (C14) in the fulmars, and by PFUnA (C11) in the murres (Butt et al., 2007). ΣPFCA levels in livers of murres and fulmars sampled from Prince Leopold Island in 2008 were about an order of magnitude lower than in samples collected in 2003–2004 (Braune et al., 2010; Butt et al., 2007). The reverse trend was seen in murre and fulmar eggs between those two time periods (Braune and Letcher, 2013). Hepatic PFOS concentrations in fulmars from our study were over an order of magnitude higher than levels found in fulmar livers sampled from the Faroe Islands in 1998–1999 (Bossi et al., 2005). Regional differences in PFAS concentrations could be attributed to a combination of factors. The transport of PFASs to remote locations may occur by direct transport via ocean currents or indirectly via atmospheric transport of volatile precursor compounds which are subsequently degraded or undergo in vivo transformation (Armitage et al., 2009; Busch et al., 2010). Disparate regions of the Arctic are influenced by air currents from different regions (Macdonald et al., 2000) which could result in the delivery of varying patterns of volatile precursors (Butt et al., 2010; Houde et al., 2011). The only quantifiable precursor compound in our study was PFOSA which was detected in one fulmar from the Minarets. The general absence of measurable concentrations of the PFAS precursors analyzed (i.e. PFOSA, N-Me-FOSA, FTOHs, FTUCAs) is consistent with the findings of Braune and Letcher (2013) for murre and fulmar eggs and supports the view that these precursor compounds are readily oxidized and degraded to PFOS and PFCAs (Ellis et al., 2004; Stock et al., 2007; Young et al., 2007; Parsons et al., 2008; Frömel and Knepper, 2010). From an oceanic perspective, surface seawater in the Canadian Archipelago and northern Hudson Bay is primarily of Pacific origin (Bidleman et al., 2007; Jones et al., 2003), whereas the Labrador Sea reflects the influx of North Atlantic waters (Bidleman et al., 2007; Jones et al., 2003; Yamashita et al., 2008) and the European Arctic is influenced primarily by Atlantic Ocean waters (Busch et al., 2010; Butt et al., 2010). Therefore, differences in PFAS patterns and concentrations noted between murres and fulmars in our study could be at least partially attributed to the influence of different source regions, transport and exposure pathways as well as differing overwintering areas utilized by these two species as discussed earlier for the HCHs. 4.4. Toxicological significance Letcher et al. (2010) recently reviewed the literature on exposure and effects of organohalogen contaminants in arctic wildlife, including seabirds, and found that there were several species and populations of concern. Those included marine birds from across the Arctic such as black-legged kittiwakes (Rissa

tridactyla) from Northern Norway, Canadian high Arctic areas and the Barents Sea, northern fulmars from the Canadian high Arctic, Alaska (Aleutian Archipelago) and Svalbard (Bjørnøya), and glaucous gulls (Larus hyperboreus) from Northern Baffin Bay and Svalbard (Bjørnøya). In our study, organochlorine concentrations found in the murre and fulmar livers were relatively low and likely not of toxicological concern. Despite their widespread occurrence in the environment, limited information is available on the toxicity of PBDEs (Ji et al., 2011). Although studies have reported PBDE-related endocrine disrupting and reproductive effects in birds at environmentally relevant concentrations (see Shaw and Kannan (2009) and references therein), we could find no toxicological threshold levels for PBDEs in avian liver. Given that PBDE levels in arctic biota are generally lower than in biota from more southerly latitudes (de Wit et al., 2010) and that the ΣPBDE concentrations found in livers of murres and fulmars in our study were not unusually high compared with other arctic seabird species (see Letcher et al., 2010), we have no reason to suspect PBDE-associated toxicological effects. However, given the lack of toxicological information available, caution is warranted. The effects of PFASs on wildlife are also not well characterized, particularly for arctic biota (Letcher et al., 2010). Although the toxicities of PFOS and PFOA have been extensively studied, there is a lack of information for many of the PFASs including the longchain PFCAs (Kannan, 2011). However, it seems that the PFCA precursors are more toxic than the PFCAs themselves (Phillips et al., 2007). Based on the toxicity threshold levels found in the literature, the low or non-detectable PFOA levels found in the murre and fulmar livers are not of toxicological concern. Laboratory studies have shown that PFOS is toxic to birds with effects ranging from decreased weight gain and increased liver mass (Newsted et al., 2005, 2006) to higher mortality, reduced hatchability and histopathological changes in the liver (Molina et al., 2006). Newsted et al. (2005) estimated a toxicity reference value (TRV) of 0.6 μg g  1 ww and a predicted no effect concentration (PNEC) of 0.35 μg g  1 ww for PFOS in liver for top avian predators. The mean hepatic PFOS concentration for the fulmars from Prince Leopold Island exceeded both the PNEC and TRV for liver. Some individual murres and fulmars from the Minarets, as well as a few murres from Prince Leopold and Akpatok Islands, also exceeded the PNEC, and four fulmars and one murre from Prince Leopold Island and two fulmars from the Minarets also exceeded the TRV for PFOS. Given these levels for PFOS, and the lack of information available for PBDEs, continued monitoring of the high Arctic colonies and, in particular, northern fulmars, is warranted.

5. Conclusions Given the variations observed for organohalogen contaminants among colonies and between murres and fulmars in this study, the reasons for inter-colony and inter-species differences in contaminant exposure are probably variable and complex. Potential factors involved include variation in source regions, contaminant transport and exposure pathways, as well as differences among colonies in their diets and overwintering areas. The differences in contaminant exposure among colonies and species observed in this study highlight the importance of continued monitoring of multiple colonies and species in the Canadian Arctic at regular intervals in the future in order to assess changing exposure patterns. As noted by Leat et al. (2013), some patterns may be best explained by variation in wintering areas and continued tracking of wintering movements should form an integral part of an ongoing monitoring strategy.

B.M. Braune et al. / Environmental Research 134 (2014) 46–56

Acknowledgments Thanks to S. Smith, P. Smith, I. Storm, S. Suppa and J. Szucs for help collecting the specimens; G. Savard, S. Robinson and students from Nunavut Arctic College for assistance with dissections; S. Lemieux and A. Idrissi of Laboratory Services, and S. Chu of the Organic Contaminants Research Laboratory at NWRC for the contaminant analyses; and C. Eberl for graphics support. Logistical support was provided by Environment Canada, Natural Resources Canada (Polar Continental Shelf Program) and the Nattivak Hunters' and Trappers' Organization, Nunavut Arctic College (Environmental Technology Program).

Appendix A. Supplementary materials Supplementary data associated with this article can be found in the online version at http://dx.doi.org/10.1016/j.envres.2014.06. 019.

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