Temporal trends of polyfluoroalkyl compounds (PFCs) in liver tissue of grey seals (Halichoerus grypus) from the Baltic Sea, 1974–2008

Chemosphere 84 (2011) 1592–1600

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Temporal trends of polyfluoroalkyl compounds (PFCs) in liver tissue of grey seals (Halichoerus grypus) from the Baltic Sea, 1974–2008 Johanna Kratzer a,b, Lutz Ahrens a,⇑, Anna Roos c, Britt-Marie Bäcklin c, Ralf Ebinghaus a a

Helmholtz-Zentrum Geesthacht, Centre for Materials and Coastal Research, Institute of Coastal Research, Department for Environmental Chemistry, D-21502 Geesthacht, Germany University of Applied Sciences Weihenstephan, Triesdorf, Germany c Department of Contaminant Research, Swedish Museum of Natural History, PO Box 50007, SE 104 05 Stockholm, Sweden b

a r t i c l e

i n f o

Article history: Received 8 February 2011 Received in revised form 20 May 2011 Accepted 23 May 2011 Available online 15 June 2011 Keywords: Temporal trend Grey seal PFCs PFASs PFOS PFOA

a b s t r a c t Temporal trends of polyfluoroalkyl compounds (PFCs) were examined in grey seal (Halichoerus grypus) liver from the Baltic Sea over a period of 35 years (1974–2008). In total, 17 of 43 PFCs were found, including the perfluoroalkyl sulfonates (C4–C10 PFSAs), perfluorooctanesulfinate (PFOSi), long chain perfluoroalkyl carboxylates (C7–C14 PFCAs), and perfluoroalkyl sulfonamides (i.e., perfluorooctane sulfonamide (FOSA) and N-ethyl perfluorooctane sulfonamide (EtFOSA)), whereas saturated and unsaturated fluorotelomer carboxylates, shorter chain PFCAs and perfluoroalkyl phosphonic acids were not detected. Perfluorooctane sulfonate (PFOS) was the predominant compound (9.57–1444 ng g 1 wet weight (ww)), followed by perfluorononanoate (PFNA, 0.47–109 ng g 1 ww). C6–C8 PFSAs, PFOSi and C7–C13 PFCAs showed statistically significant increasing concentrations between 1974 and 1997, with a peak in 1997 and then decreased or levelled off (except for C12 and C13 PFCAs). FOSA had a different temporal trend with a maximum in 1989 followed by significant decreasing concentrations until 2008. Toxicological implications for grey seals are limited, but the maximal PFOS concentration found in this study was about 40 times lower than the predicted lowest observed effect concentrations (LOEC). The statistically significant decreasing concentrations or levelling off for several PFCs in the relative closed marine ecosystem of the Baltic Sea indicate a rapidly responding to reduced emissions to the marine environment. However, the high concentrations of PFOS and continuing increasing concentrations of the longer chain PFCAs (C12– C14) shows that further work on the reduction of environmental emissions of PFCs are necessary. Ó 2011 Elsevier Ltd. All rights reserved.

1. Introduction In the past decade, polyfluoroalkyl compounds (PFCs) have gained an increasing scientific interest due to their persistence and bioaccumulative potential in the marine environment. Several recent studies detected noticeable concentrations of PFCs in marine biota from all over the world, even from remote areas far away from sources (Butt et al., 2010; Sturm and Ahrens, 2010). PFCs are stable against high temperatures, photolytic, chemical and biological degradation, and are widely used in industry and consumer products such as impregnating agents, in non-stick cookware, pesticides and aqueous film forming foams since over 50 years (Kissa, 2001). The biggest manufacturing company worldwide was the 3 M Company in the US, producing perfluorosulfonyl fluoride (POSF) based products using electrochemical fluorination (ECF). In 2000, however, the 3 M Company announced to phase-

⇑ Corresponding author. Present address: Environment Canada, Science and Technology Branch, Air Quality Research Division, 4905 Dufferin Street, Downsview, Ontario, Canada M3H 5T4. Tel.: +1 416 739 4473; fax: +1 416 739 4179. E-mail address: [email protected] (L. Ahrens). 0045-6535/$ - see front matter Ó 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2011.05.036

out the POSF-related substances, as a result of noticeable concentration in occupationally exposed persons and in the environment (3M, 2000). However, perfluorooctane sulfonate (PFOS) is still produced in Southeast Asia (Paul et al., 2009). In October 2006, the European Union (EU), formed a directive, which prohibits the general use of PFOS and their derivates from June 2008 (European Parliament and European Community Directive 2006/122/ECOF, 2006). In addition, PFOS has been added to the persistent organic pollutants (POPs) list of the Stockholm Convention in May 2009 resulting in global restrictions on its uses and production (UNEP, 2010). Being the marine top predator in the fairly enclosed Baltic Sea, the grey seal (Halichoerus grypus) population has had a history of problems connected with a high contaminant load in their feed. From approximately 100 000 grey seals in the beginning of the 1900s, the population decreased to only a few thousand animals in the 1980s (Hårding and Härkönen, 1999). The reason for the decline was hunting and diseases, including sterility among females (Helle et al., 1976). The industrial chemicals polychlorinated biphenyls (PCBs) were pointed out as main reason for the diseases, which included occluded claw deformations, uterine occlusions uteri stenosis and/or uterine occlusions, bone loss and leyomioma

J. Kratzer et al. / Chemosphere 84 (2011) 1592–1600

(Helle et al., 1976; Bergman, 1999). After PCBs were banned until the mid 1980s, concentrations in biota, including seals in the Baltic Sea, have decreased and grey seal reproductive health status has improved (Bergman, 1999). In 2008, the population estimate was at least 26 000 grey seals in the Baltic Sea (Karlsson et al., 2007). Thus, the Baltic grey seals are exposed to a variety of environmental toxins and are sensitive to contaminants. The suitable properties of PFCs under severe conditions entail their widespread usage and emphasize the relevance for them as a group of environmental pollutants which could be harmful to marine mammals like grey seals. The aim of this study was to examine temporal trends of PFCs in grey seals from the Baltic Sea between 1974 and 2008. Most investigations report data on PFOS and perfluorooctanoate (PFOA), while less is known about other PFCs in the environment. The specific objectives of this study include: (i) to determine temporal trends of PFCs in grey seal liver from the Baltic Sea over a period of 35 years (1974–2008), (ii) to investigate the occurrence and composition of 43 individual PFCs in these samples in order to identify changes over the time, and (iii) provide some toxicological implications for grey seals exposed by PFCs. 2. Materials and methods 2.1. Sample collection A total of 78 liver tissue aliquots from grey seals were obtained from the Environmental Specimen Bank (ESB) at the Swedish Museum of Natural History. The seals were collected from Baltic Sea between 1969 and 2008 (Fig. S1, Supplementary material). Grey seal samples have been collected and stored in aluminum foil and polyethylene bags under controlled conditions in the ESB ( 20 °C) and were subsampled and delivered to the HelmholtzZentrum Geesthacht. During subsampling, the surface layer of the sample, which was next to the aluminum foil, was removed, in order to prevent contamination from the foil. No cross contamination between samples is suspected, as all samples were stored in aluminum foil and then in at least two plastic bags under vacuum. All liver samples for PFC analysis were taken with stainless steel instruments, placed into polypropylene (PP) bags and stored in a 20 °C freezer until analysis. Only juvenile grey seals were included in this study, in order to exclude age as a factor of variability (Ahrens et al., 2009b). A detailed list with sampling time and location, gender, age, body measurements and cause of death can be found in Table S1 in the Supplementary material. 2.2. Extraction and analysis The target analytes included perfluoroalkyl sulfonates (PFSAs), perfluoroalkyl sulfinates (PFSiAs), perfluoroalkyl phosphonic acids (PFPAs), perfluoroalkyl carboxylates (PFCAs), fluortelomer carboxylic acids (FTCAs), fluortelomer unsaturated carboxylic acids (FTUCAs), perfluoroalkyl sulfonamides (FASAs) and perfluoroalkyl sufonamidoethanols (FASEs) (for details see Table S2 in the Supplementary material). PFCs in liver samples were extracted based on the solid–liquid extraction method described by Powley et al. (2005), Ahrens et al. (2009a). Briefly, liver samples were homogenized in polypropylene (PP) tubes using an Ultraturax disperser (T 25 basic Ultraturrax, IKA, Germany) with plastic dispersing (made of polycarbonate and polysulfone). Aliquots of 1 g of sample material were spiked with 10 ng of an internal standards (IS) mixture (Table S2, Supplementary material). Samples were extracted three times with 5 mL of acetonitrile for 30 min in an ultrasonic bath. The solvent and tissue was separated by centrifugation at

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5000 rpm for 10 min (Universal 320, Hettich Zentrifugen). The 15 mL extracts were combined and concentrated by rotary evaporation followed by gentle nitrogen blow down to 1 mL. The acetonitrile extract was further purified using a dispersive clean up with ENVI-Carb (100 mg, 1 mL, 100–400 mesh, Supelco, USA) and glacial acetic acid (Powley et al., 2005). After centrifugation the extract was concentrated by gentle nitrogen to 150 lL. Prior to injection, 50 lL (20 ng absolute) of N-deuterioethylperfluoro-1-octanesulfonamidoacetic acid (d5-EtFOSAA) was added to the extract as injection standard (InjS). Analytes were detected and quantified using a HPLC (Agilent 1100; Agilent Technologies, Palo Alto, Canada) coupled to a triple-quadrupole mass spectrometer (MS/MS; API 3000; Applied Biosystems/MDS SCIEX, Streetsville, ON, Canada) interfaced with an electrospray ionization (ESI) source in a negative-ion mode (HPLC-(–)ESI-MS/MS) as previously described (Yamashita et al., 2005). Quantification was performed based on response factors of the target compounds and their corresponding IS. The ratio of both response factors was used for recovery correction. For quantification the linear range of 0.1–300 ng mL 1 was used with a fixed IS concentration of 50 ng mL 1. As the analytical standards were not available for perfluoropentanesulfonate (PFPS), perfluorononanesulfonate (PFNS), perfluorohexadecanoate (PFPDA) and perfluoroheptadecanoate (PFHpDA), they were integrated into the method taking the MS/MS parameters of the compound having one carbon atom less in the carbon chain and their calibration was used for the quantification. Hence, the results given for PFPS, PFNS, PFPDA and PFHpDA should be considered only as an estimation. 2.3. Quality control The analytical quality of the laboratory has been approved in interlaboratory studies (Van Leeuwen et al., 2009). As standard procedure, instrumental detection limits (IDLs), method detection limits (MDLs), method blanks, matrix spiked recoveries, and matrix effects were examined (for details see Ahrens et al., 2009c). For method reproducibility and corrections, 20 mass-labelled standards were used (Table S3, Supplementary material). The InjS was used to monitor variability of the instrument response and the recovery of the IS was calculated by the related response of the InjS. Signal suppressions were observed by the factor of 0.88–0.98 caused by the matrix (for details see Ahrens et al., 2009c). The blank concentrations (n = 8) were generally <1% of the concentrations measured in the samples. The IDLs and instrumental quantification limits (IQLs) were calculated by extrapolating instrument response in a calibration standard to a concentration that would give a S/N value of 3 and 10, respectively. The MDLs and method quantification limits (MQLs) were calculated by extrapolating instrument response in liver samples to a concentration that would give a S/N value of 3 and 10, respectively. The MDLs ranged from a few pg g 1 wet weight (ww) (e.g., perfluorooctanesulfinate (PFOSi)) to a few ng g 1 ww (e.g., PFOS), depending on the compound (Table S4, Supplementary material). The matrix spike recoveries of the mass-labelled standards ranged between 57% (d5-N-ethyl perfluorooctane sulfonamidoethanol (EtFOSA)) and 91% (13C4-PFOSi). 2.4. Statistical methods The statistical analyses were performed using Microsoft Excel 2003 and SPSS for Windows 16.0 at a significance level of a = 0.05. Concentrations less than the MDL were reported as not detected, while concentrations above the MDL and below the MQL were described as
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Table 1 Geometric mean concentration (ng g

1

ww) and ranges of detected PFCs in liver tissue of grey seals from the Baltic Sea.a.

Yearb

PFBS

PFPSc

PFHxS

PFHpS

PFOS

PFDS

PFOSi

PFHpA

PFOA

PFNA

PFDA

PFUnDA

PFDoDA

PFTriDA

PFTeDA

FOSA

EtFOSA

P PFCs

MDL 1969 (n = 1) 1974 (n = 2) 1975 (n = 1) 1976 (n = 2) 1977 (n = 2) 1978 (n = 6) 1979 (n = 4) 1980 (n = 1) 1981 (n = 1) 1983 (n = 1) 1984 (n = 2) 1985 (n = 2) 1986 (n = 3) 1987 (n = 1) 1988 (n = 4) 1989 (n = 3) 1990 (n = 1) 1993 (n = 1) 1995 (n = 2) 1996 (n = 4) 1997 (n = 7) 1998 (n = 2) 1999 (n = 2) 2000 (n = 1)

0.006 1.2

0.011 nd

0.021 nd

0.022 nd

0.042 12

0.008 nd

0.001 0.01

0.005 nd

0.007 nd

0.008 0.8

0.004 0.6

0.003 0.6

0.002 0.2

0.002 0.1

0.002 nd

0.020 8.8

0.097 5.7

30

0.2 (0.1–0.2) 0.2

nd

nd

0.06 (0.1–0.1) 0.01

nd

nd nd

1.5 (1.3–1.7) 0.3

1.1 (0.9–1.2) 0.3

0.3 (0.3–0.4) 0.1

0.2 (0.2–0.2) 0.1

8.5 (6.6–11) 2.8

2.6 (1.2–5.7) nd

132

nd

2.6 (1.9–3.4) 0.6

nd

nd

115 (74–179) 24

nd

nd

0.04 (nd–0.1) nd

0.2 (0.1–0.5) 0.2 (0.1–0.2) 0.2 (nd–3.5) 0.01 (nd–0.1) nd

nd

nd

nd

nd

nd

nd

nd

nd nd

nd

nd

0.2 (0.1–0.2) 0.1 (0.1–0.2) 0.2 (0.1–0.6) 0.2 (0.1–0.6) 0.2

0.1 (0.1–0.2) 0.2 (0.1–0.2) 0.04 (nd–0.5) 0.2 (0.1–0.4) 0.2

0.004 (nd–0.2) 0.01 (nd–0.1) nd

3.8 (1.7–8.1) 13 (9.3–19) 10 (3.5–30) 17 (7.7–47) 2.9

0.5 (nd–6.9) 0.1 (nd–3.1) nd

105

nd

0.04 (nd–0.2) 0.04 (nd–0.2) nd

0.6 (0.4–0.8) 0.6 (0.4–1.0) 0.8 (0.4–2.5) 1.0 (0.4–2.8) 0.8

2.1 (1.8–2.5) nd

nd

0.6 (0.4–1.1) 0.6 (0.4–3.6) 0.7 (0.4–1.7) 0.8 (0.3–2.5) 0.5

0.01 (nd–0.1) nd

0.01 (nd–0.2) nd

0.03 (nd–0.2) 0.01 (nd–0.3) 0.01 (nd–0.3) nd

0.9 (0.5–1.9) 2.1 (1.2–3.6) 2.0 (1.1–5.1) 2.7 (1.0–6.7) 1.3

44

nd

0.01 (nd–0.1) 0.1 (0.1–0.2) 0.03 (0.01–0.05) 0.2 (0.1–0.4) nd

nd

nd

35 (9.6–129) 109 (102–117) 41 (13–157) 83 (13–157) 0.02

0.03

nd

nd

0.1

126

nd

0.04

nd

nd

8.9

2.8

2.7

0.4

0.5

0.1

9.4

nd

151

nd

nd

nd

nd

185

nd

0.05

nd

nd

13

4.2

3.9

0.6

0.7

0.2

4.1

nd

212

0.02 (nd–0.1) 0.01 (nd–0.1) 0.02 (nd–0.1) nd

nd

nd

3.2 (2.6–3.9) 3.7 (3.5–3.9) 3.3 (2.6–5.3) 11

2.9 (2.3–3.8) 4.4 (4.3–4.4) 3.8 (3.0–5.7) 9.7

0.5 (0.4–0.6) 0.8 (0.7–0.9) 0.6 (0.5–0.9) 1.2

0.7 (0.6–0.9) 0.7 (0.5–1.0) 1.0 (0.9–1.2) 1.5

0.2 (0.1–0.2) 0.2 (0.2–0.3) 0.03 (nd–0.2) 0.2

11 (6.7–17) 16 (7.8–35) 16 (11–27) 16

nd

328

1.0

9.4 (5.9–15) 12 (10–13) 12 (6.8–19) 35

333

nd

0.06 (nd–0.8) 0.02 (nd–0.1) 0.3 (nd–8.6) 0.8

nd

nd

0.1 (0.1–0.1) 0.1 (0.1–0.4) 0.2 (0.2–0.2) 0.2

164

nd

0.01 (nd–0.04) 0.03 (nd–0.2) 0.01 (nd–0.1) 0.1

nd

nd

136 (36–164) 295 (245–355) 291 (218–373) 362

nd

nd

0.05 (nd–0.2) 0.05 (nd–0.2) 0.3 (0.2–0.3) 0.2

nd

439

nd

nd

284 (206–426) 293 (198–421) 326

0.01 (nd–0.1) 0.04 (nd–0.2) 0.04

0.2 (0.1–0.3) 0.2 (0.2–0.3) 0.2

15 (7.6–23) 19 (18–22) 29

5.1 (2.4–12) 5.5 (4.0–10) 6.9

5.6 (2.9–14) 5.8 (3.9–12) 5.8

0.7 (0.4–1.6) 0.8 (0.6–1.5) 0.7

1.1 (0.8–2.2) 1.5 (1.1–3.1) 0.6

0.1 (nd–0.2) 0.01 (nd–0.3) 0.1

12 (9.1–21) 9.5 (7.8–11) 27

0.1 (nd–1.5) 0.2 (nd–1.6) nd

339

nd

0.1 (nd–6.3) 2.1 (1.7–2.7) 3.3

324

nd

0.1 (0.1–0.2) 0.2 (0.1–0.4) 0.2

nd

nd

0.6 (0.4–0.7) 1.0 (0.9–1.2) 1.0 1.1

0.2

484

nd

0.1

nd

0.1

31

9.7

9.3

1.1

1.7

0.2

11

9.1

558

2.4 (1.9–3.1) 0.9 (0.6–1.6) 1.2 (0.4–1.9) 2.6 (1.8–3.7) 1.4 (0.9–2.1) 1.1

0.7 (0.6–0.9) 0.3 (0.1–1.0) 0.6 (0.2–1.1) 1.0 (0.8–1.3) 0.4 (0.2–0.7) 0.4

620 (607–633) 429 (168–770) 457 (316–666) 825 (561–1213) 447 (399–500) 465

0.03 (nd–0.2) 0.1 (nd–0.2) 0.02 (nd–0.4) 0.3 (0.2–0.6) 0.1 (0.1–0.2) 0.2

0.2 (0.1–0.3) 0.1 (0.1–0.2) 0.2 (0.1–0.4) 0.3 (0.2–0.3) 0.3 (0.2–0.4) 0.1

0.01 (nd–0.03) nd

0.2 (nd–11) 0.3 (0.1–1.7) 6.0 (0.5–19) 11 (10–11) 2.8 (0.9–9.3) 0.2

32 (12–86) 26 (13–82) 66 (20–104) 91 (89–92) 27 (9.1–78) 15

8.5 (5.3–14) 8.0 (3.9–35) 12 (9.5–13) 23 (21–25) 7.5 (3.9–14) 4.1

9.9 (6.2–16) 9.4 (5.5–25) 12 (8.4–15) 25 (17–37) 8.0 (5.4–12) 5.9

1.3 (0.9–1.9) 1.0 (0.6–2.2) 1.3 (0.9–1.7) 3.3 (1.8–6.1) 0.9 (0.7–1.1) 0.8

3.1 (2.7–3.7) 2.0 (1.3–5.1) 2.6 (1.6–3.6) 5.8 (4.2–8.0) 2.2 (1.9–2.7) 2.8

0.02 (nd–0.3) 0.02 (nd–0.5) 0.02 (nd–0.3) 0.02 (nd–0.7) 0.01 (nd–0.3) nd

5.7 (2.8–12) 4.1 (3.3–5.2) 9.9 (3.2–23) 6.3 (3.9–10) 7.1 (5.6–9.0) 1.9

0.3 (nd–2.3) nd

685

nd

(0.03) nd

nd

nd

nd

nd

nd

0.01 (nd–0.1) 0.03 (nd–0.2) 0.1 (0.1–0.1) nd

nd nd nd

nd nd nd

nd nd

nd 0.01 (nd–0.3) nd

nd

0.005 (nd–0.1) nd 0.01 (nd–0.02) nd

nd

0.2 (nd–10) nd 0.3 (nd–1.6) nd

29

126 55

5.9

401

481 569 994 505 497

J. Kratzer et al. / Chemosphere 84 (2011) 1592–1600

0.1

nd

563

523

468

3.2 (2.1–5.3) 1.2 (0.3–12) 1.8 (1.4–2.3) 100% 0.3 (0.2–0.5) 0.3 (0.1–0.6) 0.04 (nd–0.5) 63% 3.4 (1.4–11) 3.9 (1.2–7.9) 3.5 (2.9–4.6) 97% 1.1 (0.4–5.1) 1.3 (0.5–3.1) 1.4 (1.3–1.5) 100%

3.1. PFC concentrations and component profiles

0.01 (nd–0.04) 6%

nd

nd

0.05 (nd–4.6) 0.04 (nd–3.4) 0.6 (nd–11) 62%

21 (12–38) 20 (9.3–68) 74 (45–109) 100%

6.8 (4.3–11) 7.9 (2.9–21) 17 (13–22) 100%

7.8 (3.3–26) 8.7 (2.9–23) 13 (11–15) 100%

3. Results and discussion

c

b

a

27%

nd

nd

0.03 (nd–0.3) 0.02 (nd–0.2) 0.03 (0.01–0.1) 21% nd

nd = values are < MDL. Values in brackets are between MDL and MQL. n = number of samples. To be considered as an estimation due to the lack of standard substance.

0.1 (0.1–0.1) 0.01 (nd–0.5) 0.09 (0.04–0.2) 94% 0.04 (nd–1.0) 0.04 (nd–0.5) 0.04 (nd–0.2) 44% 423 (215–1444) 479 (156–1072) 451 (398–494) 100% 0.1 (nd–0.7) 0.2 (nd–0.7) 0.3 (0.2–0.4) 76%

nd nd

0.8 (0.4–1.9) 0.7 (0.2–3.0) 0.8 (0.4–1.2) 64%

nd

nd nd

2002 (n = 3) 2003 (n = 3) 2004 (n = 1) 2005 (n = 4) 2006 (n = 8) 2008 (n = 3) % detected

1595

MDL. PFCs detected in less than 50% of the samples were excluded from the statistical comparison of means, temporal trends and correlations. Temporal trends of individual PFCs were performed using the two-way analysis of variance (ANOVA) applying natural-logarithm-transformed mean concentrations for each year to reach normality and homogeneity of the variance. The year 1969 was excluded in the regression analyses, because of the small sample number (n = 1). Pearson analysis was used for correlations between individual compound concentrations. Doubling time and half life were determined with t1/2 = ln(2)/m, where m represents the slope of the natural logarithm transformed liver concentration versus time. It should be noted that the calculation of the doubling time and half life are subject to uncertainties like differences in food supply or mother–child transfer depending on the individual animal.

0.1 (nd–1.4) 0.1 (nd–3.3) 0.1 (nd–0.7) 28%

997

702

11 (7.2–15) 2.8 (1.7–4.8) 1.0 2.8 (1.9–3.4) 4.7 (3.6–8.0) 5.2 0.9 (0.6–1.3) 1.6 (1.1–3.4) 1.7 0.02 (nd–0.2) nd nd

0.7 (0.5–0.9) 0.7 (0.3–1.2) 1.6

0.2 (0.2–0.3) 0.1 (0.01–0.3) 0.4

317 (262–437) 645 (362–975) 944

0.1 (0.05–0.3) 0.3 (0.2–0.3) 0.4

0.2 (0.1–0.4) 0.1 (0.05–0.2) 0.1

nd

0.4 (0.2–1.1) 0.4 (0.1–0.9) nd

25 (15–37) 19 (8.36–50) 14

7.0 (5.6–8.8) 11 (5.1–24) 12

7.4 (5.3–8.9) 16 (8.8–33) 16

0.04 (nd–0.3) 0.4 (0.3–0.5) 0.4

1.7 (nd–18) 0.2 (nd–1.9) nd

375

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In total, 17 of 43 target analytes were found in the liver samples, including C4–C8 and C10 PFSAs, PFOSi, C7–C14 PFCAs, perfluorooctane sulfonamide (FOSA) and EtFOSA. RPFCs concentration ranged between 29 ng g 1 ww (1975) and 1532 ng g 1 ww (2005) (geometric mean (GM): 289 ng g 1 ww, median: 385 ng g 1 ww, n = 78) (Table 1). The detection of C5 and C8 PFSA, PFOSi, C9–C13 PFCAs, FOSA and EtFOSA in 1969 indicates that the PFC contamination of the Baltic Sea area started 1969 or earlier. PFOS was the predominant compound, ranging from 9.6 ng g 1 ww (1976) to 1444 ng g 1 ww (2005) (GM: 220 ng g 1 ww, n = 78). Previous investigations reported lower concentrations of PFOS in ringed seals (Pusa hispida) (130–1100 ng g 1 ww) and grey seals (140–360 ng g 1 ww) from the Baltic Sea (Kannan et al., 2002), but higher concentrations in harbor seals (Phoca vitulina) from the German Bight (7.2–2407 ng g 1 ww) (Ahrens et al., 2009b). However, the concentration of PFOS reported here were 2–10 times higher than those found in ringed seals and grey seals from the Arctic (Giesy and Kannan, 2001). The high PFOS levels indicate the existence of a local source in the Baltic Sea area. The other PFSAs were detected in 18–74% of the samples and were about one to three orders of magnitude lower than PFOS. Interestingly, perfluorobutanesulfonate (PFBS) was only observed in the samples before 1990 with a maximum concentration of 3.5 ng g 1 ww, however, most measurements were less than 0.1 ng g 1 ww. PFBS might originate from impurities of POSFbased products, however, further studies are necessary to determine the source for PFBS in the Baltic Sea. Conversely, perfluorohexanesulfonate (PFHxS) and perfluorodecanesulfonate (PFDS) were detected the first time in 1987 and 1984, respectively and PFPS and perfluoroheptanesulfonate (PFHpS) were detected more frequently after 1984. PFOSi was found in 94% of the samples with a maximum concentration of 0.5 ng g 1 ww, which is in a similar range as those in harbor seals from the German Bight (n.d. 1.7 ng g 1 ww) (Ahrens et al., 2009b). PFOSi is proposed to be an intermediate of the biotransformation of EtFOSA to PFOS (Rhoads et al., 2008). However, the results of PFOSi must be considered carefully, because this compound can be decomposed very fast in the presence of oxygen to PFSAs. FOSA was found in all samples (GM: 6.0 ng g 1 ww) and EtFOSA was detected in 28% of the samples (GM: 0.03 ng g 1 ww). Among the PFCAs, perfluorononanoate (PFNA) was the dominant compound, followed by perfluoroundecanoate (PFUnDA) with a concentration range of 0.5–109 ng g 1 ww and 0.3–37 ng g 1 ww, respectively. The fact that PFNA was the second prevalent compound after PFOS is consistent with several previous studies

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(Martin et al., 2004; Van de Vijver et al., 2005). Differently, PFUnDA concentrations were dominating in ribbon seals (Histriophoca fasciata) from Alaska (n = 8) (Quakenbush and Citta, 2008), while PFNA was observed to be higher than PFOS in Baikal seals (Pusa sibirica) (n = 44) (Ishibashi et al., 2008a). The contribution of PFOS was in average 86% of the total PFC burden between 1969 and 2008 (Fig. 1), while the contribution of the other PFSAs (i.e., PFBS, PFPS, PFHxS, PFHpS and PFDS) and P PFOSi (classified as ‘‘others’’ in Fig. 1) was <1% of the PFC. Furthermore, the mean contribution of RPFCA to the total PFC burden was 10-fold lower than of RPFSA and increased from 4% in 1974 to 21% in 2008 (mean ± SD: 9 ± 4%). The mean contribution of PFNA and PFUnDA to RPFCAs was 53 ± 8% and 19 ± 5%, respectively. Perfluoroheptanoate (PFHpA) and PFOA were only detected sporadically which could be explained with their low bioaccumulation potential and its excretion by urine and feces without biotransformation (Kudo and Kawashima, 2003). The first detection of PFOA was in 1977 which may indicate a start of PFOA contamination in this area. An explanation of the increasing contribution of C10– C13 PFCAs between 1999 and 2004 (Fig. 1) could be increasing emissions and higher bioaccumulation potential of these compounds. Interestingly, the contribution of FOSA and EtFOSA to PFOS decreased from 1974 to 2008 (Fig. 1). A study on rainbow trout liver microsomes confirmed that EtFOSA can be biotransformed to FOSA and ultimately to PFOS (Tomy et al., 2004). However, the contribution of FOSA and EtFOSA was too small to have an influence on the change in concentration of PFOS over the time.

Most of the PFC congeners were significantly correlated to each other (Table S5, Supplementary material). This indicates that the PFCs had a common source as reported in previous studies (Ahrens et al., 2009b; Yeung et al., 2009). Overall, the pattern and generally high PFC contamination suggest distinctive sources around the Baltic Sea. They are several possible causes for the contamination of the Baltic Sea with PFCs. The rivers were considered as the main input pathway into the marine environment (McLachlan et al., 2007). Possible point sources into the rivers and directly into the Baltic Sea are municipal and industrial wastewater treatment plant (WWTP) effluents and landfill leachates (Ahrens et al., 2009a; D’eon et al., 2009; Busch et al., 2010). Other possible sources for PFCs in the Baltic Sea are, for example, dry or wet deposition from the atmosphere or surface runoff (Kwok et al., 2010; Kim and Kannan, 2007). 3.2. Temporal trends of PFSAs and PFSiAs Concentrations of PFOS increased statistically significant between 1974 and 2008 (p < 0.0001), but after 1997 the data showed a higher scattering. However, the temporal trend after 1997 is not significant. PFHxS, PFHpS and PFOSi increased significantly from 1974 to 1997, reached a peak in 1997 and then subsequently decreased significantly between 1997 and 2008 (Fig. 2 and Table 2). This represents an annual increase of 12–17% per year (doubling time: 2.5–5.6 years, 1974–1997) and annual decreased of 6.0–9.4% per year (half-live: 3.4–9.9 years, 1997–2008). PFBS, PFPS and PFDS

Fig. 1. Composition profiles of (A) PFOS, others (i.e., C4–C7 and C10 PFSAs, and PFOSi), FASAs (i.e., FOSA and EtFOSA) and (B) of individual PFCAs in grey seals from the Baltic Sea between 1969 and 2008.

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were found in 27%, 18% and 42% of the samples and were therefore not used for statistical trend analysis. In general, the maximum in 1997 and increasing and decreasing trend before and after, respectively, imply changes of the emission patterns over time. Previous temporal trend studies in marine species have shown increasing PFOS concentrations in polar bears (Ursus maritimus) until 2002 (Smithwick et al., 2006) and 2006 (Dietz et al., 2008) and Baikal seals between 1992 and 2005 (Ishibashi et al., 2008a). Decreasing PFSA concentrations were found after the mid of 1990s in harbor seals from the German Bight (Ahrens et al., 2009b) and Arctic ringed seals (Butt et al., 2007). It is interesting to note that the peak concentration, which was in 1997, was also found in previous studies (Holmström et al., 2005; Butt et al., 2007; Ahrens et al., 2009b) whereas the phase out of the POSFbased production was in 2002 (3M, 2000). This indicate that the

production and use and/or the discharge in the marine environment was reduced before the implementation of the phased out. 3.3. Temporal trends of PFCAs and FOSA Concentrations of PFOA and PFNA increased statistically significant between 1974 and 1997 (p < 0.0001) and then decreased significantly from 1997 to 2008 (p < 0.05) (Fig. 3 and Table 2), which is similar as observed for the PFSAs and PFOSi. The PFDA and PFUnDA concentrations increased also significantly from 1974 to 1997 (p < 0.0001) and levelled off after 1997. However, the levelling off after 1997 was not statistically significant. No levelling off was found for C12–C14 PFCA and PFOS after 1997, which is in contrast to the temporal trends of the PFHxS, PFHpS, PFOSi and C8–C11 PFCAs. Increasing temporal trends of C8–C11 PFCAs represent an

Fig. 2. Temporal trends of PFHxS, PFHpS, PFOS and PFOSi in grey seal livers from the Baltic Sea, 1969–2008. The plots display the geometric means (circles) and the median (green squares) together with the individual analysis (small dots), the 95% confidence intervals of the geometric means, and a seven-point running mean smoother (dashed line). (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)

Table 2 Geometric mean and 95% confidence interval (CI) of PFCs (ng g 1 ww), annual rate of change, doubling times and half-lives (in italics) (years) in the whole (1974–2008), early (1974–1997) and late (1997–2008) sampling period in grey seal liver from the Baltic Sea (n = 78).a Analyte 1974–2008 Geometric mean 95% CI Change per year (%) p Doubling times/half-lives (year)

PFHxS

PFHpS

PFOS

PFOSi

PFOA

PFNA

PFDA

PFUnDA

PFDoDA

PFTriDA

PFTeDA

FOSAb

0.20 0.12– 0.21 0.4 <0.0001 3.8

0.13 0.09– 0.18 28 <0.0001 7.2

228 163– 319 43 <0.0001 6.7

0.08 0.05– 0.11 2.2 NS NS

0.12 0.06– 0.25 16 <0.001 5.3

14 10–18

4.4 3.3–5.7

4.7 3.6–6.1

1.1 0.73–1.6

27 <0.0001 7.5

43 <0.0001 7.3

0.03 0.02– 0.06 19 <0.0001 7.2

6.0 4.7–7.6

55 <0.0001 7.1

0.70 0.57– 0.87 34 <0.0001 10

38 <0.0001 5.8

1974–1998 Change per year (%) p

12 <0.0001

16 <0.0001

171 <0.0001

17 <0.0001

7.0 <0.0001

26 <0.0001

36 <0.0001

31 <0.0001

15 <0.0001

25 <0.0001

Doubling times/half-lives (year)

2.5

4.3

4.5

5.6

2.5

4.0

4.8

4.7

6.8

4.7

0.5 NS (0.06) NS (42)

1998–2008 Change per year (%) p Doubling times/half-lives (year) a b

6.0 0.048 9.9

8.6 0.03 5.6

0.8 NS NS

9.4 0.002 3.4

13 <0.0001 1.6

4.7 0.04 14

1.2 NS NS

NS, not significant. p values based on the log-linear regression. For FOSA the early and late sampling period range from 1974–1987 and 1987–2008, respectively.

2.6 NS NS

2.8 NS NS

1.6 NS NS

55 0.02 2.3

2.1 <0.0001 18 1974– 1987 11 NS NS 1987– 2008 5.9 <0.0001 6.0

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annual increase of 7.0–36% per year from 1974 to 1997 (doubling time: 2.5–4.8 years) and annual decrease of 1.2–13% from 1997 to 2008 (half-live: 1.6 years for PFOA and 14 years for PFNA), whereas C12–C14 PFCAs increased by 19–38% between 1974 and 2008 (doubling time: 5.8–10 years) (Table 2). The most previous studies have shown increasing concentrations of PFCAs in marine mammals (Smithwick et al., 2006; Dietz et al., 2008; Sturm and Ahrens, 2010). A few studies observed a levelling off such as in Arctic ringed seals (Butt et al., 2007) and harbor seals from the German Bight (Ahrens et al., 2009b), however, these temporal trends were not statistically significant for the longer chain PFCAs. The decreasing PFCA concentrations could be caused by the reduction of PFC emissions and/or production shift to other PFCs. Recently, a global stewardship program was launched by the US Environmental Protection Agency (EPA) and eight major fluoropolymer and telomer manufacturers in 2006 to reduce PFOA and related chemicals from facility emissions and product content by 95% until 2010, and to work toward elimination of emissions and content by 2015 (USEPA, 2006). This, in turn, could result in a further reduction of the PFCA contamination in the marine mammals, however, more studies are necessary to verify this assumption.

FOSA showed a different temporal trend in comparison to the other PFCs in this study. The FOSA concentrations increased from 1974 to 1987 and then decreased statistically significant between 1987 and 2008 (p < 0.0001). This represents an annual decrease of 5.9% per years after 1989 (half-live: 6.0 years). The decreasing temporal trend of FOSA is consistent with several previous studies (Smithwick et al., 2006; Butt et al., 2007; Hart et al., 2008; Ahrens et al., 2009b). It is interesting to note that the peak concentration for FOSA was around 1989 and therefore earlier than for C6–C8 PFSAs, PFSiA, C8–C11 PFCAs. This could be caused by a production shift and/or reduction of the emissions of FOSA after 1989. Another reason might be that FOSA biodegraded partly to PFOS (Tomy et al., 2004) and therefore can be eliminated faster from the marine food web than the PFSAs and PFCAs. 3.4. Toxic implications of PFCs Toxicological studies indicate that the distribution, accumulation and toxic effects of PFCs in living organisms depend on the functional group and carbon chain length of the compound. Increasing bioaccumulation factors (BCFs) were found in biota with

Fig. 3. Temporal trends of PFCAs and FOSA in grey seal livers from the Baltic Sea, 1969–2008. The plots display the geometric means (circles) and the median (green squares) together with the individual analysis (small dots), the 95% confidence intervals of the geometric means, and a seven-point running mean smoother (dashed line). (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)

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Fig. 4. Profiles of linear and branched isomers of (A) PFOS and (B) FOSA detected in livers of grey seals from the Baltic Sea from 1969–2008.

increasing perfluoroalkyl chain length (generally, >8 carbons) (Martin et al., 2003). Moreover, PFSAs showed greater BCFs than corresponding PFCAs of the same carbon chain length (Martin et al., 2003). This is in agreement with our results showing the highest concentrations for PFOS and the longer chain PFCAs (C8– C14), while the shorter chain PFCAs (C < 7) were not detected. Toxicological implications for grey seals exposed with PFCs are not known. The potential toxicities of PFCs based on laboratory studies of monkeys and rats are reduced body weight, increased liver weight, hepatoxicity, alteration of hepatic lipid metabolism and peroxisome, reduction of serum cholesterol and thyroid hormones (Lau et al., 2004). An in vitro study on Baikal seals using a PPARa reporter gene assay determined the lowest observed effect concentrations (LOEC) with 64 lg g 1 for PFOS, PFNA and PFDA and 58 lg g 1 for PFOA and PFUnDA (Ishibashi et al., 2008b). Maximum PFOS concentrations (i.e., 1444 ng g 1 ww) in grey seals presented here were about 40 times lower than the LOEC which suggest PPARa mediated effects are unlikely. However, results from Baikal seals showed that the PPARa-CYP4A signaling pathway in the liver may be influenced by exposure to PFCs (Ishibashi et al., 2008b). This is especially noteworthy because PFOS concentrations in grey seals in this study are about 10 times higher than in Baikal seals. Ultimately, species differences can be assumed for toxicology effects in marine mammals and grey seals are exposed to a mixture of various PFCs (this study) and other organic pollutants (Helle et al., 1976; Blomkvist et al., 1992).

3.5. PFOS and FOSA isomer profiles The ECF production process of PFOS and its derivatives (e.g., FOSA) yields in a mixture of linear and branched isomers (630%), whereas the telomer-derived products from the telomerization (TM) process are strictly the linear isomer (Giesy and Kannan, 2002; De Silva and Mabury, 2006). However, little is known about the fate of branched isomers in the environment. The PFOS and FOSA isomer profiles in grey seal liver are shown in Fig. 4. Only the linear PFOS and FOSA isomer was quantified, because of the lack of calibration standards for the branched isomers. Thus, the area of the sum of branched PFOS isomers was compared with the linear isomer. Branched isomers of PFCAs in grey seal livers were not observed. Linear isomers of PFOS and FOSA dominated the composition profiles. The sum of branched PFOS isomers ranged from 7% in 1974 to 22% in 2005 (mean ± SD, 14 ± 4%). Compared to the sum of PFOS isomers, the contribution of branched isomers increased over time with highest content observed between 2004 and

2006. The sum of branched FOSA isomers increased from 2% in 1969 to 28% in 2004 (mean value 7 ± 6%). The increasing content of branched isomers over the time could be due the lower blood depuration half-lives of the branched isomers compared to the corresponding linear isomer (Benskin et al., 2009). Although the ECF production process was phased out, POSF-based products are still in use and can enter the Baltic Sea from river inputs and atmospheric deposition. However, most probably local sources are dominant due to the fact that in the Baltic Sea drainage basin live 85 million people and over 250 rivers contribute a volume of 660 km3 water per year to the Baltic Sea (HELCOM, 1993). In addition, there is only a limited water exchange with the Atlantic Ocean because the water can only flow out into the North Sea through the Danish strait. This can lead to an accumulation of chemicals in the Baltic Sea over the time. 4. Conclusion In this study temporal trends of several PFCs were evaluated in grey seals from the Baltic Sea between 1969 and 2008. Most PFCs showed statistically significant increasing concentrations between 1974 and 1997. After 1997 the PFC concentrations decreased (i.e. PFHxS, PFHpS, PFOSi, PFOA, PFNA, FOSA) or levelled off (i.e. PFDA, PFUnDA), while time trends of C12–C14 PFCAs continued to increase the whole study period. The predominant compound detected was PFOS which maximal concentration was by a factor of about 40 lower than the predicted LOEC. Additionally, the isomer profile of PFOS and FOSA in grey seal livers demonstrated an increasing content of branched isomers over time. The Baltic Sea is a relatively closed ecosystem, and is influenced by over 250 streams, which drain into the basin from industrial and high populated areas. As grey seals are living their entirely life in the Baltic Sea area, they are continuously exposed to PFCs in the aqueous phase and in their food. The main sources for PFCs in the Baltic Sea are the discharge of the rivers (McLachlan et al., 2007), WWTP effluents and landfill leachates (Ahrens et al., 2009a; D’eon et al., 2009; Busch et al., 2010) as well as diffuse sources like dry and wet atmospheric deposition (Kwok et al., 2010; Kim and Kannan, 2007). Ultimately, PFCAs and PFSAs are not biodegradable (Martin et al., 2003) and the seawater and sediment is the final sink for PFCs (Yamashita et al., 2005). Overall, our results indicate that the contamination of grey seals with PFCs decreased for the most compounds over time. However, the contamination for some PFCs still increase and further trend studies are necessary in order to evaluate future temporal trends, bioaccumulation pathways, toxicology effects and correlation of PFC concentration with gender and age.

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