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Environmental pollutants in the Swedish marine ecosystem, with special emphasis on polybrominated diphenyl ethers (PBDE)
Chemosphere 82 (2011) 1286–1292
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Environmental pollutants in the Swedish marine ecosystem, with special emphasis on polybrominated diphenyl ethers (PBDE) Pernilla Carlsson a,b,c,⇑, Dorte Herzke b, Margareta Wedborg c, Geir Wing Gabrielsen a a
Norwegian Polar Institute, Fram Centre, Hjalmar Johansens Gate 14, NO-9296 Tromsø, Norway Norwegian Institute for Air Research, Fram Centre, Hjalmar Johansens Gate 14, NO-9296 Tromsø, Norway c Department of Chemistry, University of Gothenburg, SE-412 96 Gothenburg, Sweden b
a r t i c l e
i n f o
Article history: Received 19 May 2010 Received in revised form 11 November 2010 Accepted 7 December 2010 Available online 8 January 2011 Key words: PBDE PCB DDT PFC Common eider Herring gull
a b s t r a c t Levels of polybrominated diphenyl ethers (PBDEs), polybrominated biphenyls (PBBs), polychlorinated biphenyls (PCBs), dichlorodiphenyltrichloroethane (DDT) and perfluorinated organic compounds (PFCs) were analysed in whole herring (Clupea harengus) and sprat (Sprattus sprattus), eggs from common eider (Somateria mollissima) and eggs and livers from herring gull (Larus argentatus) from the Swedish west coast. The contaminant values obtained were compared with published values from the Arctic marine ecosystem. Tetra- and penta-brominated PBDEs were detected at low levels in herring, sprat and common eider (RPBDE 0.3–2.0 ng g1 ww), while the levels were higher in the herring gull samples (RPBDE 1.3–29.9 ng g1 ww). Hexa-decaBDEs were also found in samples from herring gulls. Eggs from herring gulls from the sub-Arctic contained four times more PBDE than the Swedish herring gulls eggs. Fish samples from the Arctic had two times higher levels of PBDEs and DDTs than similar samples from Sweden. The higher levels of contaminants in fish and seabirds from the Arctic reflect differences in transport processes, feeding ecology (reflected by trophic levels) and metabolism. PBDEs contributed to <10% of the total contaminant load in all investigated samples. The relative contribution of DDTs was higher in fish and bird samples from the Arctic when compared to Swedish samples, e.g. 65% in glaucous gull livers compared to 10% in herring gull livers. This study shows that even though the Swedish west coast is more urban than the Arctic, higher pollutants levels are found in seabird species from the Arctic. Ó 2010 Elsevier Ltd. All rights reserved.
1. Introduction Trend data of polybrominated diphenyl ethers (PBDEs) from Swedish biota show decreasing and quite low levels in marine biota, even though the Swedish west coast and its surroundings have had some point sources of PBDEs (de Wit, 2002). Levels of BDE-47, -99 and -100 in eggs from common guillemot (Uria aalge) from Stora Karlsö, Baltic Sea, increased during the early 1970s, peaked during the late 1980s and significantly decreased during the 1990s (Sellström et al., 2003). Decreasing concentrations of BDE-47 have also been observed in herring (Clupea harengus) and cod (Gadus morhua) from the Swedish west coast. The levels of polychlorinated biphenyls (PCBs) in herring from the Baltic Sea have decreased by 70–90% since the seventies, while dichlorodiphenyltrichloroethane (DDT) has decreased by 10–20% in cod liver and herring muscles (Bignert et al., 2008). Nevertheless, levels in top predators like peregrine falcons (Falco peregrinus) are still
⇑ Corresponding author. Present address: UNIS, Pb. 156, NO-9171 Longyearbyen, Norway. Tel.: +47 79023315; fax: +47 7902 3301. E-mail address: [email protected] (P. Carlsson). 0045-6535/$ - see front matter Ó 2010 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2010.12.029
high (>1000 ng g1 lipid weight (lw)) (Sellström, 1999; Johansson et al., 2009). The levels of PBDEs in Arctic birds have been low (de Wit et al., 2003). However, recent studies show an increase in biota, lake sediment and air (de Wit et al., 2003; de Wit et al., 2006; Evenset et al., 2007; Manø et al., 2008). Ocean currents, winds and rivers are important transport pathways of contaminants to the Arctic (de Wit et al., 2003). The predominant long-range transport of POPs to the Arctic lead consequently to higher concentration levels in the Arctic environment compared to contamination levels in low latitude regions (AMAP, 2009). All investigated pollutants in the present study have shown to undergo atmospheric long-range transport into the Arctic, and especially the congeners with lowest vapour pressure (Manø et al., 2008; AMAP, 2009; Hung et al., 2009). Studies from north Canada indicate increasing concentrations of PBDEs in ivory gulls (Pagophila eburnea) (Braune et al., 2007). This trend has also been observed among glaucous gulls (Larus hyperboreus) from Bjørnøya (Knudsen et al., 2005). Similar to other lipophilic substances, PBDEs tend to bioaccumulate very easily, especially the lower brominated congeners (de Wit, 2002). Nevertheless, decaBDE (BDE-209) has been found in polar cod (Boreogadus saida), at approximately the same levels
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as the more common pentabrominated congeners BDE-99 and 100 (Sørmo et al., 2006). Atmospheric and ocean transport of pollutants as well as the feeding ecology of different species affect the pollutant levels in animals. Arctic seabirds have a higher metabolism per gram body weight compared to species from temperate areas (Ellis and Gabrielsen, 2002). As a result, seabirds from the Arctic have a higher food intake, which may result in a higher intake of pollutants in the Arctic seabirds. As an example, four times higher levels of PBDEs have been found in glaucous gulls than in polar bears from Svalbard (Verreault et al., 2005). Investigations of pollutants from the Swedish west coast are scarce. The main objective of the present study was to investigate the levels of environmental pollution in biota from the Swedish west coast, and to compare these values with literature data of levels of pollutants in fish and seabird species from the Arctic. 2. Materials and methods 2.1. Sampling procedure Herring and sprat (Sprattus sprattus) were caught in Skagerrak during August and September 2008. Eggs from common eiders (Somateria mollissima) and herring gulls (Larus argentatus) were collected in May 2008 in the Strömstad archipelago (58°N, 11°E), Swedish west coast (Fig. 1). Adult herring gulls were shot during autumn 2008 in the same area. All samples were kept frozen until analysis. Pools of whole fish (three fish in each pool, n = 9, three
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pools of sprats and six of herrings), egg content (n = 10 for each species) and six livers from adult herring gulls were analysed. The Swedish fish were a homogenous group with little differences in concentrations and are hereafter referred to as ‘‘fish’’ instead of the specific species if nothing else is stated. All samples were collected in accordance with the regulations issued by the Swedish Animal Research Authority. 2.2. Sample treatment 2.2.1. Persistent organic pollutants (POPs) Tissue and egg samples were extracted and prepared as described previously (Vetter et al., 2007; Herzke et al., 2009). Briefly, the samples were homogenised, subsequently dried in a tenfold amount of dry sodium sulphate, and extracted after addition of the 13C-labelled internal standards (IS) with cyclohexane/acetone (50:50 v/v). The amount of extractable organic material was determined gravimetrically. Lipids were removed with a gel permeation chromatography (GPC) system. An additional fractionation was carried out on a Florisil™ column. Before quantification, octachloronaphtalene was added to all samples as a recovery standard. For quantification of all compounds, reference materials were obtained from Cambridge Isotope Laboratories (Woburn, MA, USA). The following 13C-labelled compounds were used as ISs, representing each group of analytes: 13C-p,p0 -dichlorodiphenyldichloroethene (DDE); 13C-p,p0 -DDT; 13C-labelled PCBs 28, 52, 101, 118, 138, 153 and 180; 13C-labelled BDE-28, 47, 99, 153, 183 and 209. The other chemicals used were of pesticide grade (Merck, Darmstadt, Germany). The results were not corrected for recovery. 2.2.2. PFCs Prior to analysis, samples were extracted and prepared as described previously (Herzke et al., 2009) with some modifications given here: A sample amount of 1 g of the tissue or egg was spiked with 20 ng of the IS 13C-labelled PFOS and PFOA (Wellington Laboratories Inc., USA). After extraction with 7 mL acetonitrile in an ultrasonic bath for 30 min, samples were centrifuged (10 min, 5000 rpm), and 1 mL of the supernatant solution was added to 25 mg ENVI-Carb and 50 lL glacial acetic acid and vortexed thoroughly. After additional centrifugation at 10 000 rpm for 10 min, 0.5 mL of each sample was transferred to vials. Prior to quantification, 3,7-dimethyl-branched perfluorodecanoic acid (bPFDcA, 97%, ABCR; Germany) as a recovery standard and 0.5 mL of a 2 mM aqueous ammonium acetate solution were added to all samples. 2.3. Chromatographic separation and quantification
Fig. 1. Location of sampling sites in Sweden (Swedish west coast), northern Norway (Røst and Hornøya) and Svalbard). Picture from the Norwegian Polar Institute.
2.3.1. POPs All samples were analysed for a suite of conventional POPs: CB28, 52, 99, 101, 105, 118, 138, 153, 170, 180, 183, 187, 194; p,p0 DDT, o,p0 -DDT, p,p0 -DDE and o,p0 -DDE; BB 15, 49, 52, 101, 153; BDE-28, 47, 49, 66, 71, 77, 85, 99, 100, 119, 138, 153, 154, 183, 196, 206, 207, 208, 209. For the brominated compounds an Agilent 7890A gas chromatograph (GC) was equipped with a 15 m DB5-MS column (0.25 mm id and 0.10 lm film thickness; J&W, Folsom, USA). Helium (6.0 quality, Hydrogas, Porsgrunn, Norway) was used as carrier gas at a flow rate of 1 mL min1. One microliter of the sample extract was injected with a S/SL injector at 250 °C (Agilent Technologies, 7683B). The following temperature program was used: start at 100 °C; a 20 °C min1 increase to 250 °C; a 1.5 °C min1 increase to 260 °C; a 20 °C min1 increase to 320 °C; hold 4 min. The instrument was operated in electronic ionisation (EI) mode. Detection was carried out by multiple reaction monitoring (MRM) measurements using a Quattro micro™ mass spectrometer (MS) (Micromass MS Technologies, Manchester, UK) with an
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ionisation energy of 70 eV. The mass transitions are presented in Supplementary Table 1. The transfer line temperature was held at 280 °C and the source temperature was set to 220 °C. The PCBs and DDTs were analysed with the method described by Herzke et al. (2009). 2.3.2. PFCs We analysed the ionic PFCs perfluorocarboxylates (PFCAs) with a carbon chain containing 4–13 carbons (PFBA, PFPA, PFHxA, PFHpA, PFOA, PFNA, PFDcA, PFUnA, PFDoA, PFTrA), as well as the following PFSAs; perfluorobutane (PFBS), perfluorohexane (PFHxS), perfluorooctane (PFOS), and perfluorodecane sulfonate (PFDcS) in addition to perfluorooctane sulfonamide (PFOSA). All standards were purchased from Aldrich, Fluka or ABCR with purities greater than 95%. Aliquots (50 lL) were injected on a high-performance liquid chromatograph (HPLC) (Agilent 1100; Agilent Technologies, Palo Alto, CA) coupled to electronic spray ionisation time-of-flight high resolution MS in the negative ion mode (HPLC–ESI–ToF–MS) (LCT, Micromass, Manchester, England). Compounds were separated on an ACE-C18 column (150 2.1 mm, 3 lm particle size) (ACT, Aberdeen, UK) using a gradient of 200 lL min1 methanol and water (both with 2 mM NH4OAc). The initial mobile phase condition was 50:50 methanol/water, followed by a 5 min ramp increase to 85:15, a 5 min hold at 85:15, a 0.5 min ramp to 99:1, and hold until reverting to initial condition after 15 min. Full scan (m/z 165–720) high resolution mass spectra were monitored throughout the chromatograms. An alternating cone voltage of 20 V (for PFCAs) and 40 V (for PFSAs) was applied. Further details are given in Haukås et al., 2007. 2.3.3. Quality control For each POP compound, a quantification mass and a qualifier mass were acquired relating the area proportions to the measurements of standard compounds. For PFCs, the high resolution full scan spectra, acquired for each sample, were used to control positive detections. Quality assurance of the analytical method was carried out by measurements of laboratory blanks and standard reference materials (SRM). Because of the lack of commercial SRMs for PFC a pre-analysed fish tissue sample with origin from the EU project PERFORCE was used with every batch of 10 samples. One sample of fish tissue was analysed using both IS (labelled and non-labelled) in triplicate. The results for PFOS and PFOA varied between 5% and 10% using the different IS. A laboratory blank was analysed for POPs and PFCs with every sample batch. No lab contamination for any of the analysed components was detected. The LOD was calculated as three times the signal-to-noise ratio for each compound and sample type and is provided as Supplementary Table 2. As a consequence, the detection limits depended on the sample amount extracted and varied between samples as well as between analytes (0.02–0.77 ng g1 for PBDEs and PBBs, 0.01–0.22 ng g1 for PCBs and DDTs and 0.01–1.20 ng g1 for PFCs,). The limit of quantification (LOQ) was set to 10 times laboratory blank for all target analytes. The recoveries were 66–74% for PBDEs, 56% for PCBs, 65% for DDTs and 95% for the PFCs. 3. Results and discussion The mean concentrations of the PBDE congeners are presented in Table 3, together with the sum of PCBs, DDTs and PFCs. BDE47 was the only congener that was found in all samples. It was also the most abundant congener, followed by BDE-99 and -100 (Fig. 2). The RPBDE contributed to 3–10% of the total burden of the analysed substances (Fig. 3), which is in line with a previous study
which concluded that PCBs and DDE contributed to 90% of the total burden, while RPBDE contributed to less than 2% (Herzke et al., 2003). PBBs were only detected in a few samples and will therefore not be discussed in detail. Almost all of the investigated congeners of PCB were detected in all samples. O,p0 -DDE was not detected, but p,p0 -DDE was found in all samples. The levels of PFCs were quite similar in the eggs and in the liver samples from seabirds. The herring gull samples contained the highest levels of all analysed substances among the Swedish samples. To our knowledge, no investigations have previously been conducted on PBDEs in liver samples from herring gulls collected at the Swedish west coast. 3.1. Swedish data 3.1.1. Fish BDE-47 had the highest concentrations at similar levels as earlier reported in cod muscles from Swedish waters (Bignert et al., 2008). However, earlier reported levels in herring muscles from the Swedish west coast were 15 times higher than the levels recorded in the present study (Darnerud et al., 2001). These variations could be explained by temporal differences in combination with differences in the type of tissue analysed. The fish samples in the present study had the lowest levels of PCBs and DDTs of the samples studied (Table 3). The fact that these fish occupy a lower trophic level than the seabirds is probably the most important reason for these lower pollutant levels (Borgå et al., 2001). The level of RDDT was in accordance with earlier reported values in whole herring from the same area (Bignert et al., 2008). The levels of RPFCs were the lowest in the present study (0.6– 3.5 ng g1 ww) compared to earlier reported levels of up to 14 ng g1 ww of PFOSA in liver samples from herring collected off the Swedish west coast (Bignert et al., 2008). However, livers generally contain higher levels of pollutants than muscle samples. 3.1.2. Seabirds The only congener detected in all egg samples from common eider was BDE-47 (Table 3). The PBDE levels were similar to levels in eider eggs from Sklinna, Norway (Herzke et al., 2009). This indicates that neither of the eider populations is close to the source of pollutants and they feed at a low trophic level. BDE-47 dominated the herring gull eggs and liver samples, followed by BDE-99 and -100, (Fig. 2) in accordance with other egg studies, e.g. ivory gulls (Braune et al., 2007), herring gulls (Norstrom et al., 2002) and glaucous gulls (Knudsen et al., 2005). Several other congeners were also found in the herring gull liver samples (Table 3). One of the herring gull eggs contained more decaBDE than BDE47, which is remarkable since low brominated congeners are more likely than higher brominated congeners to be transferred from the hen to the eggs (Verreault et al., 2006a). DecaBDE was also detected in two of the livers, both of which had a relatively high total concentration of PBDEs (100 and 625 ng g1 lw of RPBDE). The herring gull eggs contained three times more RDDT and four times more RPCB, than eggs from the common eiders, and slightly higher levels than the herring gull livers. The p,p0 -DDE contributed to 92% and 100% of the RDDT in the common eider and the herring gull eggs, respectively. Since no DDT was present in the herring gull eggs but p,p0 -DDE was, it is likely to assume that there are differences in feeding habits, but also in metabolism and maternal energetic demands among these species (Verreault et al., 2006b; Gabrielsen, 2009). The levels of RPCB and RDDT were six and three times higher, respectively, in the eggs from Swedish eiders compared to eggs from eiders from Sklinna, Norway (Herzke et al., 2009). The Norwegian eiders probably live further away from PCB and DDT sources.
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Table 3 Concentrations (ng g1 ww) of the substances analysed and detected >LOQ. PFOS was detected in all of the samples, PFHxS and PFHpS and some of the PFCAs were found in the bird samples, but not in the fish samples. ce = common eider, hg = herring gull. Substance
% Lipids PBDEs BDE-28 BDE-47 BDE-49 BDE-66 BDE-71 BDE-99 BDE-100 BDE-153 BDE-154 BDE-183 BDE-209 RPBDE RPBB RPCB RDDT RPFC
Fishes n = 9
ce egg n = 10
Mean
Range
14%
10.2–21.9
0.05 0.59 0.18
0.22 0.21
1.4 11.4 3.8 8.0
SD
0.01 0.2 0.1
0.1 0.1 0.01
Mean
Range
18%
14.7–30.0
0.28
0.34
0.22
0.4
0.6
4 2 5
108.4 20.1 91.1
n.d. 0.15–0.55 n.d. n.d. n.d.
hg egg n = 10 SD
0.1
Mean
Range
6%
4.4–9.6
2.58 0.11
0.2
0.4
1.39 1.42 0.61 0.36 0.58 1.3 8.3
23 8 73
205.4 19.5 29.9
0.1 0.1
hg liver n = 6 SD
1.5
Mean
Range
5%
3.8–5.0
2.34
0.04 0.9 1.1 0.4 0.2 0.3 1.9 6
2.65 1.02 0.73 0.46
7.5
94 9 13
84.8 13.2 70.0
SD
2
5 1 1 0.3 0.1 1 11 39 8 54
Fig. 2. Relative contribution of the different congeners to RPBDE. Congeners that were only detected in a few samples are not included in this figure.
Fig. 3. Relative distribution of PBDEs, PCBs and DDTs. Herring gull = hg, glaucous gull = gg. The PFCs were not included since they do not bioaccumulate in the same way as PBDEs, etc. do. All data concerning glaucous gulls are literature data (Verreault et al., 2007b; Verboven et al., 2008, Verreault, unpublished data).
The eggs from eiders had levels of RPFCs which were similar to those of the herring gull eggs and livers (Table 3). Common eider eggs from Sklinna, Norway, contained 25 ng g1 ww of PFCs, which is comparable to the levels recorded in the eggs in the present study (Herzke et al., 2009). The samples were dominated by PFOS, which is in accordance with other studies where livers from black guillemot (Cepphus grylle) and glaucous gull (Haukås et al., 2007) and eggs from ivory gulls (Miljeteig et al., 2009) have been analysed. Perfluorocarboxylates (PFCA) showed a similar pattern in both seabird species. Higher amounts of long chained PFCA were detected in the eider eggs. The demersal feeding habits among common eiders will expose them more to sediment bound pollutants than is the case with the herring gull (Gabrielsen, 2009). Long chain PFCs have a higher surface affinity than the short chain congeners, and are therefore associated to particles rather than to the water body. The higher variety of PFCs detected in eggs from eider might also be due to feeding closer to point sources, such as discharge from sewage treatment plants, during the period of egg formation. The common eider is a precocial species. This hatching strategy relies on major lipid and protein reserves in the eggs enabling a successful formation of the fetus (Steen and Gabrielsen,
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1986). Thus differences in mother-egg-transfer mechanisms might also explain some of the difference in PFC pattern between common eider and herring gulls.
3.2. Comparison of POPs in Sweden and the Arctic Literature data of PBDEs in herring gull egg from northern Norway (Hornøya; 70°N, 31°E and Røst; 67°N, 12°E, collected in 2003) (Helgason et al., 2009) and PBDE, PCB and DDT data from glaucous gull liver (collected during early summer 2002 in Ny-Ålesund, 79°N, 19°E) and glaucous gull egg yolk (spring 2006, Bjørnøya, 74°N, 19°E) (Verreault et al., 2007a,b; Verboven et al., 2008) were used in the comparison between Sweden and the European Arctic. The Swedish fish samples (zooplanktivorous) were compared with polar cods, since this is a zooplanktivorous high Arctic species (Hop et al., 2002; Bignert et al., 2008). Although the comparison of fish and bird data from Sweden with corresponding data from the Arctic is not ideal, the data are reasonably comparable since they were collected at approximately the same time of the year. A recent study reported slightly lower PBDE levels in shore shrimps (Palaemon adspersus and P. elegans) from Skagerrak than in ice amphipods (Gammarus wilkitzkii) from Svalbard (Sørmo et al., 2009). This is in accordance with the present study, where the pollutant burden is higher in the Arctic than in the Swedish samples (Fig. 4). RDDT dominated in the glaucous gulls, while RPCB dominated the Swedish samples, followed by RDDT and RPBDE (Fig. 2). This implies differences in especially feeding habits, but also in metabolism and pollutant sources for the species. The levels of PFCs were higher in glaucous gull eggs than in the Swedish seabird eggs.
3.2.1. Fish The congener pattern among the Swedish fish was similar to studies from the Arctic where BDE-47 dominated in polar cods. Even though polar cods live further away from urban areas, previous studies have reported twice as high levels of PBDEs and DDTs compared to the fish in the present study (Wolkers et al., 2004; Haukås et al., 2007; Tomy et al., 2008). This indicates different pollutant sources compared to the polar cod collected in the Arctic.
Fig. 4. Swedish samples compared to Norwegian herring gulls. Mean concentrations (ng g1 lw) in the Swedish and the Arctic samples. N-herring gull = herring gulls from northern Norway. Swedish samples: RPBDE: sum of BDE-28, -47, -49, 66, -71, -77, -85, -99, -100, -119, -138, -153, -154, -183, -196, -206, -207, -208 and 209. RPCB: sum of CB-28, -52, -99, -101, -118, -138, -153, -180, -183, -187, -194. RDDT: sum of p,p0 -DDT, o,p0 -DDT, p,p0 -DDE. Northern Norway samples: RPBDE: sum of BDE-28, -47, -99, -100, -153 and -154. PCBs and DDTs were not analysed in the eggs from northern Norway. Arctic data from Helgason et al. (2009).
There are few data related to the trophic level of herring and interpretations should be made carefully. However, the present results may indicate that these species have been exposed to different pollutant sources. 3.2.2. Seabirds Herring gull eggs from northern Norway contained four times as much RPBDE (Fig. 4) as the Swedish herring gull eggs and nine times higher levels of BDE-47 (Helgason et al., 2009). The differences between the two colonies indicate different pollutant sources and exposures. Hornøya and Røst are remote areas, while the Swedish west coast is more densely populated. Hence, it is remarkable that the herring gull eggs from northern Norway contain the highest levels of PBDEs. As a comparison, levels of RPBDE were 5 and 70 times higher in glaucous gull eggs and livers, respectively, compared to the same tissue samples from Swedish herring gulls (Verreault et al., 2007a,b). The levels of RPCBs and RDDTs were 5 and 39 times higher respectively, in the glaucous gull egg yolks (Verreault et al., 2007a) compared to the herring gull eggs. The difference is even larger when it comes to the seabird livers, 180 and 1170 times higher levels of RPCBs and RDDTs, respectively, in the glaucous gulls (Verreault, unpublished data). Diet has been reported as an important factor for explaining differences in PCB and PBDE levels between different populations of glaucous gulls. It has been shown that glaucous gulls feed on eggs of other marine birds which would increase their trophic level above the herring gull and explain some of the elevated findings (Bustnes et al., 2000; Steffen et al., 2006). Haukås et al. (2007) reported 39.8 ng g1 ww of PFCs in liver samples from glaucous gulls, which is higher than the levels in the herring gull livers in the present study (22.9 ng g1 ww). Since the glaucous gulls live further away from the pollution sources than the herring gulls, it was unexpected to find such large difference in contaminant levels. Differences in trophic positions were expected to compensate for this, but this is not the case in the present study. Again, feeding on eggs of other marine bird species might explain the differences observed. 3.3. Feeding ecology The disparities in pollutant levels found in different seabird species can partly be explained by different trophic levels: 4.2 and 3.4 for glaucous gulls and herring gulls, respectively (Hop et al., 2002; Ruus et al., 2002). From a recent investigation it was concluded that the trophic position was more important than the geographic distribution concerning POP levels (Steffen et al., 2006). The rate of biomagnification is higher among seabirds than in fish (Borgå, 2002). A combination of feeding at higher trophic level, having a high energy expenditure and thus a high basal metabolic rate (BMR) and food intake may result in high contaminant levels in glaucous gulls (Braune and Norstrom, 1989). Arctic seabird species have 40–50% higher BMR than temperate seabird species (Ellis and Gabrielsen, 2002). The result of a higher BMR is a higher daily energy expenditure resulting in higher food intake which in turn causes a higher intake of pollutants. Glaucous gulls are twice as heavy as herring gulls, but the reported BMR values for glaucous gulls (0.0177–0.0260 kJ g1 h1) are almost within the same range as BMR in herring gulls (0.0173–0.0193 kJ g1 h1) (Ellis and Gabrielsen, 2002). This implies that the energy expenditure and, as a consequence, the food/contaminant intake may be twice as high in glaucous gulls as compared to herring gulls due to the higher body mass of the glaucous gulls. These simple calculations may indicate that the glaucous gulls are more exposed to POPs than the herring gulls because of the higher need of nutrition/feed intake.
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4. Conclusions The Arctic has until recently, often been considered as a pristine area with low levels of pollutants compared to more urban areas. However, the present study indicates the opposite. The highest levels of all of the POPs analysed were found in the herring gulls from the west coast of Sweden. However, RPBDE, RPCB and RDDT levels reported in the literature for glaucous gulls were 70, 180 and 1170 times higher than the levels found in liver samples from the Swedish herring gulls. Arctic literature data for fish and bird eggs were in general higher than the levels in the present study. Differences in trophic levels and energy expenditure may be a more important factor than the geographical distribution. The pristine Arctic area seems to have a higher input of PBDE compared to the more densely populated Swedish west coast. It is therefore highly recommended that further studies will be carried out in order to determine if the observed increasing trend of PBDEs in the Arctic compared to more urban areas is real. Acknowledgements This study was financed by the Norwegian Polar Institute. The analyses were performed at NILU, Tromsø. We also want to acknowledge Matti Åhlund, Jan Karlsson and Birgitta Krischansson for collecting the samples, Jonathan Verreault for providing Arctic data from glaucous gulls and Roland Kallenborn and Mark Hermanson for valuable comments. Appendix A. Supplementary material Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.chemosphere.2010.12.029. References AMAP, 2009. Arctic Pollution 2009. Arctic Monitoring and Assessment Programme, Oslo, Norway, pp. xvi+88. Bignert, A., Danielsson, S., Strandmark, A., Nyberg, E., Asplund, L., Eriksson, U. et al., 2008. Sakrapport Metaller och Organiska Miljögifter i Marin Biota, Trend-och Områdesövervakning. Swedish Museum of Natural History, Department of Contaminant Research, Stockholm, p. 142. Borgå, K., 2002. Organochlorine Contaminants in Arctic Marine Food Webs: Distribution in Pelagic and Sympagic Fauna. PhD-Thesis, University of Tromsö, Tromsö. Borgå, K., Gabrielsen, G.W., Skaare, J.U., 2001. Biomagnification of organochlorines along a Barents sea food chain. Environ. Pollut. 113, 187–198. Braune, B.M., Norstrom, R.J., 1989. Dynamics of organochlorine compounds in herring gulls – 3. Tissue distribution and bioaccumulation in Lake Ontario Gulls. Environ. Toxicol. Chem. 8, 957–968. Braune, B.M., Mallory, M.L., Grant Gilchrist, H., Letcher, R.J., Drouillard, K.G., 2007. Levels and trends of organochlorines and brominated flame retardants in Ivory Gull eggs from the Canadian Arctic, 1976 to 2004. Sci. Total Environ. 378, 403– 417. Bustnes, J.O., Erikstad, K.E., Bakken, V., Mehlum, F., Skaare, J.U., 2000. Feeding ecology and the concentration of organochlorines in glaucous gulls. Ecotoxicology 9, 179–186. Darnerud, P.O., Eriksen, G.S., Jóhannesson, T., Larsen, P.B., Viluksela, M., 2001. Polybrominated diphenyl ethers: occurrence, dietary exposure, and toxicology. Environ. Health Persp. 109, 49–68. de Wit, C.A., 2002. An overview of brominated flame retardants in the environment. Chemospere 46, 583–624. de Wit, C.A., Fisk, A., Hobbs, K., Muir, D., Gabrielsen, G., Kallenborn, R., Krahn, M., Norstrom, R., Skaare, J., 2003. Persistent organic pollutants. In: Wilson, S.J., Murray, J.L., Huntington, H.P., (Eds.), AMAP II Assessment Report. Arctic Pollution Issues, Arctic Monitoring and Assessment Program. Oslo, Norway, p. +310. de Wit, C.A., Alaee, M., Muir, D.C.G., 2006. Levels and trends of brominated flame retardants in the Arctic. Chemosphere 64, 209–233. Ellis, H.I., Gabrielsen, G.W., 2002. Energetics of free-ranging seabirds. In: Marine Birds. CRC Press, Boca Raton, Florida, USA, 2002. Evenset, A., Christensen, G.N., Carroll, J., Zaborska, A., Berger, U., Herzke, D., et al., 2007. Historical trends in persistent organic pollutants and metals recorded in sediment from Lake Ellasjøen, Bjørnøya, Norwegian Arctic. Environ. Pollut. 146, 196–205.
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Geographical distribution (2000) and temporal trends (1981–2000) of brominated diphenyl ethers in Great Lakes Herring Gull eggs. Environ. Sci. Technol. 36, 4783–4789. Ruus, A., Inne Ugland, K., Utne Skaare, J., 2002. Influence of trophic position on organochlorine concentrations and compositional patterns in a marine food web. Environ. Toxicol. Chem. 21, 2356–2364. Sellström, U., 1999. Determination of Some Polybrominated Flame Retardants in Biota, Sediment and Sewage Sludge. Department of Environmental Chemistry, Institute of Applied Environmental Research (ITM), Laboratory for Anaytical Environmental Chemistry. PhD-Thesis, Stockholm University, Stockholm. Sellström, U., Bignert, A., Kierkegaard, A., Haggberg, L., de Wit, C.A., Olsson, M., Jansson, B., 2003. Temporal trend studies on tetra- and pentabrominated diphenyl ethers and hexabromocyclododecane in Guillemot egg from the Baltic Sea. Environ. Sci. Technol. 37, 5496–5501. Steen, J.B., Gabrielsen, G.W., 1986. Thermogenesis in newly hatched Eider (Somateria mollissima) and long-tailed duck (Clangula hyemalis) ducklings and Barnacle Goose (Branta leucopsis) goslings. Polar Res. 4, 181–186. Steffen, C., Borga, K., Skaare, J.U., Bustnes, J.O., 2006. The occurrence of organochlorines in marine avian top predators along a latitudinal gradient. Environ. Sci. Technol. 40, 5139–5146. Sørmo, E.G., Salmer, M.P., Jenssen, B.M., Hop, H., Bæk, K., Kovacs, K.M., et al., 2006. Biomagnification of polybrominated diphenyl ether and hexabromocyclododecane flame retardants in the polar bear food chain in Svalbard, Norway. Environ. Toxicol. Chem. 25, 2502–2511. Sørmo, E.G., Jenssen, B.M., Lie, E., Skaare, J.U., 2009. Brominated flame retardants in aquatic organisms from the North Sea in comparison with biota from the high Arctic marine environment. Environ. Toxicol. Chem. 28, 2082–2090. Tomy, G.T., Pleskach, K., Oswald, T., Halldorson, T., Helm, P.A., MacInnis, G., et al., 2008. 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Environmental pollutants in the Swedish marine ecosystem, with special emphasis on polybrominated diphenyl ethers (PBDE) Pernilla Carlsson a,b,c,⇑, Dorte Herzke b, Margareta Wedborg c, Geir Wing Gabrielsen a a
Norwegian Polar Institute, Fram Centre, Hjalmar Johansens Gate 14, NO-9296 Tromsø, Norway Norwegian Institute for Air Research, Fram Centre, Hjalmar Johansens Gate 14, NO-9296 Tromsø, Norway c Department of Chemistry, University of Gothenburg, SE-412 96 Gothenburg, Sweden b
a r t i c l e
i n f o
Article history: Received 19 May 2010 Received in revised form 11 November 2010 Accepted 7 December 2010 Available online 8 January 2011 Key words: PBDE PCB DDT PFC Common eider Herring gull
a b s t r a c t Levels of polybrominated diphenyl ethers (PBDEs), polybrominated biphenyls (PBBs), polychlorinated biphenyls (PCBs), dichlorodiphenyltrichloroethane (DDT) and perfluorinated organic compounds (PFCs) were analysed in whole herring (Clupea harengus) and sprat (Sprattus sprattus), eggs from common eider (Somateria mollissima) and eggs and livers from herring gull (Larus argentatus) from the Swedish west coast. The contaminant values obtained were compared with published values from the Arctic marine ecosystem. Tetra- and penta-brominated PBDEs were detected at low levels in herring, sprat and common eider (RPBDE 0.3–2.0 ng g1 ww), while the levels were higher in the herring gull samples (RPBDE 1.3–29.9 ng g1 ww). Hexa-decaBDEs were also found in samples from herring gulls. Eggs from herring gulls from the sub-Arctic contained four times more PBDE than the Swedish herring gulls eggs. Fish samples from the Arctic had two times higher levels of PBDEs and DDTs than similar samples from Sweden. The higher levels of contaminants in fish and seabirds from the Arctic reflect differences in transport processes, feeding ecology (reflected by trophic levels) and metabolism. PBDEs contributed to <10% of the total contaminant load in all investigated samples. The relative contribution of DDTs was higher in fish and bird samples from the Arctic when compared to Swedish samples, e.g. 65% in glaucous gull livers compared to 10% in herring gull livers. This study shows that even though the Swedish west coast is more urban than the Arctic, higher pollutants levels are found in seabird species from the Arctic. Ó 2010 Elsevier Ltd. All rights reserved.
1. Introduction Trend data of polybrominated diphenyl ethers (PBDEs) from Swedish biota show decreasing and quite low levels in marine biota, even though the Swedish west coast and its surroundings have had some point sources of PBDEs (de Wit, 2002). Levels of BDE-47, -99 and -100 in eggs from common guillemot (Uria aalge) from Stora Karlsö, Baltic Sea, increased during the early 1970s, peaked during the late 1980s and significantly decreased during the 1990s (Sellström et al., 2003). Decreasing concentrations of BDE-47 have also been observed in herring (Clupea harengus) and cod (Gadus morhua) from the Swedish west coast. The levels of polychlorinated biphenyls (PCBs) in herring from the Baltic Sea have decreased by 70–90% since the seventies, while dichlorodiphenyltrichloroethane (DDT) has decreased by 10–20% in cod liver and herring muscles (Bignert et al., 2008). Nevertheless, levels in top predators like peregrine falcons (Falco peregrinus) are still
⇑ Corresponding author. Present address: UNIS, Pb. 156, NO-9171 Longyearbyen, Norway. Tel.: +47 79023315; fax: +47 7902 3301. E-mail address: [email protected] (P. Carlsson). 0045-6535/$ - see front matter Ó 2010 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2010.12.029
high (>1000 ng g1 lipid weight (lw)) (Sellström, 1999; Johansson et al., 2009). The levels of PBDEs in Arctic birds have been low (de Wit et al., 2003). However, recent studies show an increase in biota, lake sediment and air (de Wit et al., 2003; de Wit et al., 2006; Evenset et al., 2007; Manø et al., 2008). Ocean currents, winds and rivers are important transport pathways of contaminants to the Arctic (de Wit et al., 2003). The predominant long-range transport of POPs to the Arctic lead consequently to higher concentration levels in the Arctic environment compared to contamination levels in low latitude regions (AMAP, 2009). All investigated pollutants in the present study have shown to undergo atmospheric long-range transport into the Arctic, and especially the congeners with lowest vapour pressure (Manø et al., 2008; AMAP, 2009; Hung et al., 2009). Studies from north Canada indicate increasing concentrations of PBDEs in ivory gulls (Pagophila eburnea) (Braune et al., 2007). This trend has also been observed among glaucous gulls (Larus hyperboreus) from Bjørnøya (Knudsen et al., 2005). Similar to other lipophilic substances, PBDEs tend to bioaccumulate very easily, especially the lower brominated congeners (de Wit, 2002). Nevertheless, decaBDE (BDE-209) has been found in polar cod (Boreogadus saida), at approximately the same levels
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as the more common pentabrominated congeners BDE-99 and 100 (Sørmo et al., 2006). Atmospheric and ocean transport of pollutants as well as the feeding ecology of different species affect the pollutant levels in animals. Arctic seabirds have a higher metabolism per gram body weight compared to species from temperate areas (Ellis and Gabrielsen, 2002). As a result, seabirds from the Arctic have a higher food intake, which may result in a higher intake of pollutants in the Arctic seabirds. As an example, four times higher levels of PBDEs have been found in glaucous gulls than in polar bears from Svalbard (Verreault et al., 2005). Investigations of pollutants from the Swedish west coast are scarce. The main objective of the present study was to investigate the levels of environmental pollution in biota from the Swedish west coast, and to compare these values with literature data of levels of pollutants in fish and seabird species from the Arctic. 2. Materials and methods 2.1. Sampling procedure Herring and sprat (Sprattus sprattus) were caught in Skagerrak during August and September 2008. Eggs from common eiders (Somateria mollissima) and herring gulls (Larus argentatus) were collected in May 2008 in the Strömstad archipelago (58°N, 11°E), Swedish west coast (Fig. 1). Adult herring gulls were shot during autumn 2008 in the same area. All samples were kept frozen until analysis. Pools of whole fish (three fish in each pool, n = 9, three
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pools of sprats and six of herrings), egg content (n = 10 for each species) and six livers from adult herring gulls were analysed. The Swedish fish were a homogenous group with little differences in concentrations and are hereafter referred to as ‘‘fish’’ instead of the specific species if nothing else is stated. All samples were collected in accordance with the regulations issued by the Swedish Animal Research Authority. 2.2. Sample treatment 2.2.1. Persistent organic pollutants (POPs) Tissue and egg samples were extracted and prepared as described previously (Vetter et al., 2007; Herzke et al., 2009). Briefly, the samples were homogenised, subsequently dried in a tenfold amount of dry sodium sulphate, and extracted after addition of the 13C-labelled internal standards (IS) with cyclohexane/acetone (50:50 v/v). The amount of extractable organic material was determined gravimetrically. Lipids were removed with a gel permeation chromatography (GPC) system. An additional fractionation was carried out on a Florisil™ column. Before quantification, octachloronaphtalene was added to all samples as a recovery standard. For quantification of all compounds, reference materials were obtained from Cambridge Isotope Laboratories (Woburn, MA, USA). The following 13C-labelled compounds were used as ISs, representing each group of analytes: 13C-p,p0 -dichlorodiphenyldichloroethene (DDE); 13C-p,p0 -DDT; 13C-labelled PCBs 28, 52, 101, 118, 138, 153 and 180; 13C-labelled BDE-28, 47, 99, 153, 183 and 209. The other chemicals used were of pesticide grade (Merck, Darmstadt, Germany). The results were not corrected for recovery. 2.2.2. PFCs Prior to analysis, samples were extracted and prepared as described previously (Herzke et al., 2009) with some modifications given here: A sample amount of 1 g of the tissue or egg was spiked with 20 ng of the IS 13C-labelled PFOS and PFOA (Wellington Laboratories Inc., USA). After extraction with 7 mL acetonitrile in an ultrasonic bath for 30 min, samples were centrifuged (10 min, 5000 rpm), and 1 mL of the supernatant solution was added to 25 mg ENVI-Carb and 50 lL glacial acetic acid and vortexed thoroughly. After additional centrifugation at 10 000 rpm for 10 min, 0.5 mL of each sample was transferred to vials. Prior to quantification, 3,7-dimethyl-branched perfluorodecanoic acid (bPFDcA, 97%, ABCR; Germany) as a recovery standard and 0.5 mL of a 2 mM aqueous ammonium acetate solution were added to all samples. 2.3. Chromatographic separation and quantification
Fig. 1. Location of sampling sites in Sweden (Swedish west coast), northern Norway (Røst and Hornøya) and Svalbard). Picture from the Norwegian Polar Institute.
2.3.1. POPs All samples were analysed for a suite of conventional POPs: CB28, 52, 99, 101, 105, 118, 138, 153, 170, 180, 183, 187, 194; p,p0 DDT, o,p0 -DDT, p,p0 -DDE and o,p0 -DDE; BB 15, 49, 52, 101, 153; BDE-28, 47, 49, 66, 71, 77, 85, 99, 100, 119, 138, 153, 154, 183, 196, 206, 207, 208, 209. For the brominated compounds an Agilent 7890A gas chromatograph (GC) was equipped with a 15 m DB5-MS column (0.25 mm id and 0.10 lm film thickness; J&W, Folsom, USA). Helium (6.0 quality, Hydrogas, Porsgrunn, Norway) was used as carrier gas at a flow rate of 1 mL min1. One microliter of the sample extract was injected with a S/SL injector at 250 °C (Agilent Technologies, 7683B). The following temperature program was used: start at 100 °C; a 20 °C min1 increase to 250 °C; a 1.5 °C min1 increase to 260 °C; a 20 °C min1 increase to 320 °C; hold 4 min. The instrument was operated in electronic ionisation (EI) mode. Detection was carried out by multiple reaction monitoring (MRM) measurements using a Quattro micro™ mass spectrometer (MS) (Micromass MS Technologies, Manchester, UK) with an
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ionisation energy of 70 eV. The mass transitions are presented in Supplementary Table 1. The transfer line temperature was held at 280 °C and the source temperature was set to 220 °C. The PCBs and DDTs were analysed with the method described by Herzke et al. (2009). 2.3.2. PFCs We analysed the ionic PFCs perfluorocarboxylates (PFCAs) with a carbon chain containing 4–13 carbons (PFBA, PFPA, PFHxA, PFHpA, PFOA, PFNA, PFDcA, PFUnA, PFDoA, PFTrA), as well as the following PFSAs; perfluorobutane (PFBS), perfluorohexane (PFHxS), perfluorooctane (PFOS), and perfluorodecane sulfonate (PFDcS) in addition to perfluorooctane sulfonamide (PFOSA). All standards were purchased from Aldrich, Fluka or ABCR with purities greater than 95%. Aliquots (50 lL) were injected on a high-performance liquid chromatograph (HPLC) (Agilent 1100; Agilent Technologies, Palo Alto, CA) coupled to electronic spray ionisation time-of-flight high resolution MS in the negative ion mode (HPLC–ESI–ToF–MS) (LCT, Micromass, Manchester, England). Compounds were separated on an ACE-C18 column (150 2.1 mm, 3 lm particle size) (ACT, Aberdeen, UK) using a gradient of 200 lL min1 methanol and water (both with 2 mM NH4OAc). The initial mobile phase condition was 50:50 methanol/water, followed by a 5 min ramp increase to 85:15, a 5 min hold at 85:15, a 0.5 min ramp to 99:1, and hold until reverting to initial condition after 15 min. Full scan (m/z 165–720) high resolution mass spectra were monitored throughout the chromatograms. An alternating cone voltage of 20 V (for PFCAs) and 40 V (for PFSAs) was applied. Further details are given in Haukås et al., 2007. 2.3.3. Quality control For each POP compound, a quantification mass and a qualifier mass were acquired relating the area proportions to the measurements of standard compounds. For PFCs, the high resolution full scan spectra, acquired for each sample, were used to control positive detections. Quality assurance of the analytical method was carried out by measurements of laboratory blanks and standard reference materials (SRM). Because of the lack of commercial SRMs for PFC a pre-analysed fish tissue sample with origin from the EU project PERFORCE was used with every batch of 10 samples. One sample of fish tissue was analysed using both IS (labelled and non-labelled) in triplicate. The results for PFOS and PFOA varied between 5% and 10% using the different IS. A laboratory blank was analysed for POPs and PFCs with every sample batch. No lab contamination for any of the analysed components was detected. The LOD was calculated as three times the signal-to-noise ratio for each compound and sample type and is provided as Supplementary Table 2. As a consequence, the detection limits depended on the sample amount extracted and varied between samples as well as between analytes (0.02–0.77 ng g1 for PBDEs and PBBs, 0.01–0.22 ng g1 for PCBs and DDTs and 0.01–1.20 ng g1 for PFCs,). The limit of quantification (LOQ) was set to 10 times laboratory blank for all target analytes. The recoveries were 66–74% for PBDEs, 56% for PCBs, 65% for DDTs and 95% for the PFCs. 3. Results and discussion The mean concentrations of the PBDE congeners are presented in Table 3, together with the sum of PCBs, DDTs and PFCs. BDE47 was the only congener that was found in all samples. It was also the most abundant congener, followed by BDE-99 and -100 (Fig. 2). The RPBDE contributed to 3–10% of the total burden of the analysed substances (Fig. 3), which is in line with a previous study
which concluded that PCBs and DDE contributed to 90% of the total burden, while RPBDE contributed to less than 2% (Herzke et al., 2003). PBBs were only detected in a few samples and will therefore not be discussed in detail. Almost all of the investigated congeners of PCB were detected in all samples. O,p0 -DDE was not detected, but p,p0 -DDE was found in all samples. The levels of PFCs were quite similar in the eggs and in the liver samples from seabirds. The herring gull samples contained the highest levels of all analysed substances among the Swedish samples. To our knowledge, no investigations have previously been conducted on PBDEs in liver samples from herring gulls collected at the Swedish west coast. 3.1. Swedish data 3.1.1. Fish BDE-47 had the highest concentrations at similar levels as earlier reported in cod muscles from Swedish waters (Bignert et al., 2008). However, earlier reported levels in herring muscles from the Swedish west coast were 15 times higher than the levels recorded in the present study (Darnerud et al., 2001). These variations could be explained by temporal differences in combination with differences in the type of tissue analysed. The fish samples in the present study had the lowest levels of PCBs and DDTs of the samples studied (Table 3). The fact that these fish occupy a lower trophic level than the seabirds is probably the most important reason for these lower pollutant levels (Borgå et al., 2001). The level of RDDT was in accordance with earlier reported values in whole herring from the same area (Bignert et al., 2008). The levels of RPFCs were the lowest in the present study (0.6– 3.5 ng g1 ww) compared to earlier reported levels of up to 14 ng g1 ww of PFOSA in liver samples from herring collected off the Swedish west coast (Bignert et al., 2008). However, livers generally contain higher levels of pollutants than muscle samples. 3.1.2. Seabirds The only congener detected in all egg samples from common eider was BDE-47 (Table 3). The PBDE levels were similar to levels in eider eggs from Sklinna, Norway (Herzke et al., 2009). This indicates that neither of the eider populations is close to the source of pollutants and they feed at a low trophic level. BDE-47 dominated the herring gull eggs and liver samples, followed by BDE-99 and -100, (Fig. 2) in accordance with other egg studies, e.g. ivory gulls (Braune et al., 2007), herring gulls (Norstrom et al., 2002) and glaucous gulls (Knudsen et al., 2005). Several other congeners were also found in the herring gull liver samples (Table 3). One of the herring gull eggs contained more decaBDE than BDE47, which is remarkable since low brominated congeners are more likely than higher brominated congeners to be transferred from the hen to the eggs (Verreault et al., 2006a). DecaBDE was also detected in two of the livers, both of which had a relatively high total concentration of PBDEs (100 and 625 ng g1 lw of RPBDE). The herring gull eggs contained three times more RDDT and four times more RPCB, than eggs from the common eiders, and slightly higher levels than the herring gull livers. The p,p0 -DDE contributed to 92% and 100% of the RDDT in the common eider and the herring gull eggs, respectively. Since no DDT was present in the herring gull eggs but p,p0 -DDE was, it is likely to assume that there are differences in feeding habits, but also in metabolism and maternal energetic demands among these species (Verreault et al., 2006b; Gabrielsen, 2009). The levels of RPCB and RDDT were six and three times higher, respectively, in the eggs from Swedish eiders compared to eggs from eiders from Sklinna, Norway (Herzke et al., 2009). The Norwegian eiders probably live further away from PCB and DDT sources.
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Table 3 Concentrations (ng g1 ww) of the substances analysed and detected >LOQ. PFOS was detected in all of the samples, PFHxS and PFHpS and some of the PFCAs were found in the bird samples, but not in the fish samples. ce = common eider, hg = herring gull. Substance
% Lipids PBDEs BDE-28 BDE-47 BDE-49 BDE-66 BDE-71 BDE-99 BDE-100 BDE-153 BDE-154 BDE-183 BDE-209 RPBDE RPBB RPCB RDDT RPFC
Fishes n = 9
ce egg n = 10
Mean
Range
14%
10.2–21.9
0.05 0.59 0.18
0.22 0.21
1.4 11.4 3.8 8.0
SD
0.01 0.2 0.1
0.1 0.1 0.01
Mean
Range
18%
14.7–30.0
0.28
0.34
0.22
0.4
0.6
4 2 5
108.4 20.1 91.1
n.d. 0.15–0.55 n.d. n.d. n.d.
hg egg n = 10 SD
0.1
Mean
Range
6%
4.4–9.6
2.58 0.11
0.2
0.4
1.39 1.42 0.61 0.36 0.58 1.3 8.3
23 8 73
205.4 19.5 29.9
0.1 0.1
hg liver n = 6 SD
1.5
Mean
Range
5%
3.8–5.0
2.34
0.04 0.9 1.1 0.4 0.2 0.3 1.9 6
2.65 1.02 0.73 0.46
7.5
94 9 13
84.8 13.2 70.0
SD
2
5 1 1 0.3 0.1 1 11 39 8 54
Fig. 2. Relative contribution of the different congeners to RPBDE. Congeners that were only detected in a few samples are not included in this figure.
Fig. 3. Relative distribution of PBDEs, PCBs and DDTs. Herring gull = hg, glaucous gull = gg. The PFCs were not included since they do not bioaccumulate in the same way as PBDEs, etc. do. All data concerning glaucous gulls are literature data (Verreault et al., 2007b; Verboven et al., 2008, Verreault, unpublished data).
The eggs from eiders had levels of RPFCs which were similar to those of the herring gull eggs and livers (Table 3). Common eider eggs from Sklinna, Norway, contained 25 ng g1 ww of PFCs, which is comparable to the levels recorded in the eggs in the present study (Herzke et al., 2009). The samples were dominated by PFOS, which is in accordance with other studies where livers from black guillemot (Cepphus grylle) and glaucous gull (Haukås et al., 2007) and eggs from ivory gulls (Miljeteig et al., 2009) have been analysed. Perfluorocarboxylates (PFCA) showed a similar pattern in both seabird species. Higher amounts of long chained PFCA were detected in the eider eggs. The demersal feeding habits among common eiders will expose them more to sediment bound pollutants than is the case with the herring gull (Gabrielsen, 2009). Long chain PFCs have a higher surface affinity than the short chain congeners, and are therefore associated to particles rather than to the water body. The higher variety of PFCs detected in eggs from eider might also be due to feeding closer to point sources, such as discharge from sewage treatment plants, during the period of egg formation. The common eider is a precocial species. This hatching strategy relies on major lipid and protein reserves in the eggs enabling a successful formation of the fetus (Steen and Gabrielsen,
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1986). Thus differences in mother-egg-transfer mechanisms might also explain some of the difference in PFC pattern between common eider and herring gulls.
3.2. Comparison of POPs in Sweden and the Arctic Literature data of PBDEs in herring gull egg from northern Norway (Hornøya; 70°N, 31°E and Røst; 67°N, 12°E, collected in 2003) (Helgason et al., 2009) and PBDE, PCB and DDT data from glaucous gull liver (collected during early summer 2002 in Ny-Ålesund, 79°N, 19°E) and glaucous gull egg yolk (spring 2006, Bjørnøya, 74°N, 19°E) (Verreault et al., 2007a,b; Verboven et al., 2008) were used in the comparison between Sweden and the European Arctic. The Swedish fish samples (zooplanktivorous) were compared with polar cods, since this is a zooplanktivorous high Arctic species (Hop et al., 2002; Bignert et al., 2008). Although the comparison of fish and bird data from Sweden with corresponding data from the Arctic is not ideal, the data are reasonably comparable since they were collected at approximately the same time of the year. A recent study reported slightly lower PBDE levels in shore shrimps (Palaemon adspersus and P. elegans) from Skagerrak than in ice amphipods (Gammarus wilkitzkii) from Svalbard (Sørmo et al., 2009). This is in accordance with the present study, where the pollutant burden is higher in the Arctic than in the Swedish samples (Fig. 4). RDDT dominated in the glaucous gulls, while RPCB dominated the Swedish samples, followed by RDDT and RPBDE (Fig. 2). This implies differences in especially feeding habits, but also in metabolism and pollutant sources for the species. The levels of PFCs were higher in glaucous gull eggs than in the Swedish seabird eggs.
3.2.1. Fish The congener pattern among the Swedish fish was similar to studies from the Arctic where BDE-47 dominated in polar cods. Even though polar cods live further away from urban areas, previous studies have reported twice as high levels of PBDEs and DDTs compared to the fish in the present study (Wolkers et al., 2004; Haukås et al., 2007; Tomy et al., 2008). This indicates different pollutant sources compared to the polar cod collected in the Arctic.
Fig. 4. Swedish samples compared to Norwegian herring gulls. Mean concentrations (ng g1 lw) in the Swedish and the Arctic samples. N-herring gull = herring gulls from northern Norway. Swedish samples: RPBDE: sum of BDE-28, -47, -49, 66, -71, -77, -85, -99, -100, -119, -138, -153, -154, -183, -196, -206, -207, -208 and 209. RPCB: sum of CB-28, -52, -99, -101, -118, -138, -153, -180, -183, -187, -194. RDDT: sum of p,p0 -DDT, o,p0 -DDT, p,p0 -DDE. Northern Norway samples: RPBDE: sum of BDE-28, -47, -99, -100, -153 and -154. PCBs and DDTs were not analysed in the eggs from northern Norway. Arctic data from Helgason et al. (2009).
There are few data related to the trophic level of herring and interpretations should be made carefully. However, the present results may indicate that these species have been exposed to different pollutant sources. 3.2.2. Seabirds Herring gull eggs from northern Norway contained four times as much RPBDE (Fig. 4) as the Swedish herring gull eggs and nine times higher levels of BDE-47 (Helgason et al., 2009). The differences between the two colonies indicate different pollutant sources and exposures. Hornøya and Røst are remote areas, while the Swedish west coast is more densely populated. Hence, it is remarkable that the herring gull eggs from northern Norway contain the highest levels of PBDEs. As a comparison, levels of RPBDE were 5 and 70 times higher in glaucous gull eggs and livers, respectively, compared to the same tissue samples from Swedish herring gulls (Verreault et al., 2007a,b). The levels of RPCBs and RDDTs were 5 and 39 times higher respectively, in the glaucous gull egg yolks (Verreault et al., 2007a) compared to the herring gull eggs. The difference is even larger when it comes to the seabird livers, 180 and 1170 times higher levels of RPCBs and RDDTs, respectively, in the glaucous gulls (Verreault, unpublished data). Diet has been reported as an important factor for explaining differences in PCB and PBDE levels between different populations of glaucous gulls. It has been shown that glaucous gulls feed on eggs of other marine birds which would increase their trophic level above the herring gull and explain some of the elevated findings (Bustnes et al., 2000; Steffen et al., 2006). Haukås et al. (2007) reported 39.8 ng g1 ww of PFCs in liver samples from glaucous gulls, which is higher than the levels in the herring gull livers in the present study (22.9 ng g1 ww). Since the glaucous gulls live further away from the pollution sources than the herring gulls, it was unexpected to find such large difference in contaminant levels. Differences in trophic positions were expected to compensate for this, but this is not the case in the present study. Again, feeding on eggs of other marine bird species might explain the differences observed. 3.3. Feeding ecology The disparities in pollutant levels found in different seabird species can partly be explained by different trophic levels: 4.2 and 3.4 for glaucous gulls and herring gulls, respectively (Hop et al., 2002; Ruus et al., 2002). From a recent investigation it was concluded that the trophic position was more important than the geographic distribution concerning POP levels (Steffen et al., 2006). The rate of biomagnification is higher among seabirds than in fish (Borgå, 2002). A combination of feeding at higher trophic level, having a high energy expenditure and thus a high basal metabolic rate (BMR) and food intake may result in high contaminant levels in glaucous gulls (Braune and Norstrom, 1989). Arctic seabird species have 40–50% higher BMR than temperate seabird species (Ellis and Gabrielsen, 2002). The result of a higher BMR is a higher daily energy expenditure resulting in higher food intake which in turn causes a higher intake of pollutants. Glaucous gulls are twice as heavy as herring gulls, but the reported BMR values for glaucous gulls (0.0177–0.0260 kJ g1 h1) are almost within the same range as BMR in herring gulls (0.0173–0.0193 kJ g1 h1) (Ellis and Gabrielsen, 2002). This implies that the energy expenditure and, as a consequence, the food/contaminant intake may be twice as high in glaucous gulls as compared to herring gulls due to the higher body mass of the glaucous gulls. These simple calculations may indicate that the glaucous gulls are more exposed to POPs than the herring gulls because of the higher need of nutrition/feed intake.
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4. Conclusions The Arctic has until recently, often been considered as a pristine area with low levels of pollutants compared to more urban areas. However, the present study indicates the opposite. The highest levels of all of the POPs analysed were found in the herring gulls from the west coast of Sweden. However, RPBDE, RPCB and RDDT levels reported in the literature for glaucous gulls were 70, 180 and 1170 times higher than the levels found in liver samples from the Swedish herring gulls. Arctic literature data for fish and bird eggs were in general higher than the levels in the present study. Differences in trophic levels and energy expenditure may be a more important factor than the geographical distribution. The pristine Arctic area seems to have a higher input of PBDE compared to the more densely populated Swedish west coast. It is therefore highly recommended that further studies will be carried out in order to determine if the observed increasing trend of PBDEs in the Arctic compared to more urban areas is real. Acknowledgements This study was financed by the Norwegian Polar Institute. The analyses were performed at NILU, Tromsø. We also want to acknowledge Matti Åhlund, Jan Karlsson and Birgitta Krischansson for collecting the samples, Jonathan Verreault for providing Arctic data from glaucous gulls and Roland Kallenborn and Mark Hermanson for valuable comments. Appendix A. Supplementary material Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.chemosphere.2010.12.029. References AMAP, 2009. Arctic Pollution 2009. Arctic Monitoring and Assessment Programme, Oslo, Norway, pp. xvi+88. Bignert, A., Danielsson, S., Strandmark, A., Nyberg, E., Asplund, L., Eriksson, U. et al., 2008. Sakrapport Metaller och Organiska Miljögifter i Marin Biota, Trend-och Områdesövervakning. Swedish Museum of Natural History, Department of Contaminant Research, Stockholm, p. 142. Borgå, K., 2002. Organochlorine Contaminants in Arctic Marine Food Webs: Distribution in Pelagic and Sympagic Fauna. PhD-Thesis, University of Tromsö, Tromsö. Borgå, K., Gabrielsen, G.W., Skaare, J.U., 2001. Biomagnification of organochlorines along a Barents sea food chain. Environ. Pollut. 113, 187–198. Braune, B.M., Norstrom, R.J., 1989. Dynamics of organochlorine compounds in herring gulls – 3. Tissue distribution and bioaccumulation in Lake Ontario Gulls. Environ. Toxicol. Chem. 8, 957–968. Braune, B.M., Mallory, M.L., Grant Gilchrist, H., Letcher, R.J., Drouillard, K.G., 2007. Levels and trends of organochlorines and brominated flame retardants in Ivory Gull eggs from the Canadian Arctic, 1976 to 2004. Sci. Total Environ. 378, 403– 417. Bustnes, J.O., Erikstad, K.E., Bakken, V., Mehlum, F., Skaare, J.U., 2000. Feeding ecology and the concentration of organochlorines in glaucous gulls. Ecotoxicology 9, 179–186. Darnerud, P.O., Eriksen, G.S., Jóhannesson, T., Larsen, P.B., Viluksela, M., 2001. Polybrominated diphenyl ethers: occurrence, dietary exposure, and toxicology. Environ. Health Persp. 109, 49–68. de Wit, C.A., 2002. An overview of brominated flame retardants in the environment. Chemospere 46, 583–624. de Wit, C.A., Fisk, A., Hobbs, K., Muir, D., Gabrielsen, G., Kallenborn, R., Krahn, M., Norstrom, R., Skaare, J., 2003. Persistent organic pollutants. In: Wilson, S.J., Murray, J.L., Huntington, H.P., (Eds.), AMAP II Assessment Report. Arctic Pollution Issues, Arctic Monitoring and Assessment Program. Oslo, Norway, p. +310. de Wit, C.A., Alaee, M., Muir, D.C.G., 2006. Levels and trends of brominated flame retardants in the Arctic. Chemosphere 64, 209–233. Ellis, H.I., Gabrielsen, G.W., 2002. Energetics of free-ranging seabirds. In: Marine Birds. CRC Press, Boca Raton, Florida, USA, 2002. Evenset, A., Christensen, G.N., Carroll, J., Zaborska, A., Berger, U., Herzke, D., et al., 2007. Historical trends in persistent organic pollutants and metals recorded in sediment from Lake Ellasjøen, Bjørnøya, Norwegian Arctic. Environ. Pollut. 146, 196–205.
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