Comparison of PFASs contamination in the freshwater and terrestrial environments by analysis of eggs from osprey (Pandion haliaetus), tawny owl (Strix aluco), and common kestrel (Falco tinnunculus)

Environmental Research 149 (2016) 40–47

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Comparison of PFASs contamination in the freshwater and terrestrial environments by analysis of eggs from osprey (Pandion haliaetus), tawny owl (Strix aluco), and common kestrel (Falco tinnunculus) Ulrika Eriksson a,n, Anna Roos b, Ylva Lind b, Kjell Hope a, Alf Ekblad a, Anna Kärrman a a b

Man-Technology-Environment (MTM) Research Centre, School of Science and Technology, Örebro University, SE-701 82 Örebro, Sweden Swedish Museum of Natural History, P.O. Box 50007, SE-104 05 Stockholm, Sweden

art ic l e i nf o

a b s t r a c t

Article history: Received 9 March 2016 Received in revised form 26 April 2016 Accepted 27 April 2016

The level of PFAS (per- and polyfluorinated alkyl substances) contamination in freshwater and terrestrial Swedish environments in 2013/2014 was assessed by analyzing a range of perfluorinated alkyl acids, fluorotelomer acids, sulfonamides, sulfonamidoethanols and polyfluoralkyl phosphate diesters (diPAPs) in predator bird eggs. Stable isotopes (13C and 15N) were analyzed to elucidate the dietary source. The tawny owl (Strix aluco, n¼ 10) and common kestrel (Falco tinnunculus, n¼ 40), two terrestrial species, and the osprey (Pandion haliaetus, n¼30), a freshwater specie were included. In addition, a temporal trend (1997–2001, 2008–2009, 2013) in osprey was studied as well. The PFAS profile was dominated by perfluorooctane sulfonic acid (PFOS) in eggs from osprey and tawny owl, while for common kestrel perfluorinated carboxylic acids (∑PFCA) exceeded the level of PFOS. PFOS concentration in osprey eggs remained at the same level between 1997 and 2001 and 2013. For the long-chained PFCAs, there were a significant increase in concentrations in osprey eggs between 1997 and 2001 and 2008–2009. The levels of PFOS and PFCAs were about 10 and five times higher, respectively, in osprey compared to tawny owl and common kestrel. Evidence of direct exposure from PFCA precursor compounds to birds in both freshwater and terrestrial environment was observed. Low levels of diPAPs were detected in a few samples of osprey ( o0.02–2.4 ng/g) and common kestrel ( o0.02–0.16 ng/g) eggs, and 6:2 FTSA was detected in a majority of the osprey eggs ( o 6.3–52 ng/g). One saturated telomer acid (7:3 FTCA), which is a transformation marker from precursor exposure, was detected in all species ( o0.24–2.7 ng/g). The 15 N data showed higher levels in osprey eggs compared to tawny owl and common kestrel, indicating that they feed on a 2–3 times higher trophic level. We conclude that ospreys are continuously exposed to PFAS at levels where adverse toxic effects have been observed in birds. & 2016 Elsevier Inc. All rights reserved.

Keywords: diPAPs PFAS Bird of prey Freshwater Terrestrial

1. Introduction Per-and polyfluorinated alkyl substances (PFASs) are widely used in a large range of industrial and consumer products, despite increasing awareness and knowledge amongst scientists and policy makers about their environmental impact and fate (Buck et al., 2011; Lindstrom et al., 2011). The concerns raised around the year 2000 regarding their toxicity, persistence, and global spread led to restrictions and voluntary phase-out of the most emergent PFASs. Perfluorooctane sulfonic acid (PFOS) was phased out by the major producer 3M Company, and a few years later the voluntary stewardship program was initiated by US EPA which aimed to eliminate long-chained perfluorocarboxylic acids (PFCAs) (3M n

Corresponding author. E-mail address: [email protected] (U. Eriksson).

http://dx.doi.org/10.1016/j.envres.2016.04.038 0013-9351/& 2016 Elsevier Inc. All rights reserved.

Company, 2000; US EPA, 2006). Concurrent with decreased manufacturing and use of the aforementioned PFASs, production of shorter-chained homologues, fluorotelomer compounds, for instance polyfluoroalkyl phosphate esters (PAPs), as well as new groups of PFASs such as perfluoropolyethers (PFPEs) have increased (Wang et al., 2013). Also, a geographical shift has occurred, with decreasing PFOS production in Europe and North America and increasing PFAS production in Asia, comprising continuous production of PFOS (Wang et al., 2014a). PFOS and PFCAs (C9-C14) are persistent, resulting in bioaccumulation and biomagnification in aquatic and terrestrial food webs (Kelly et al., 2009; Tomy et al., 2009; Muller et al., 2011; Xu et al., 2014). Therefore, birds of prey are suitable for environmental biomonitoring. Since the phase-out of PFOS, temporal trend studies have shown various trends of PFASs in wild birds. For example, some studies of birds of prey have shown decreasing PFOS levels (Ahrens et al., 2011), or a peak around 2000 followed by a

U. Eriksson et al. / Environmental Research 149 (2016) 40–47

decline (Verreault et al., 2007; Holmstrom et al., 2010; Miller et al., 2015), while other studies have observed unaltered PFOS levels (Braune and Letcher, 2013; Miller et al., 2015). The variation observed between studies is likely due to their difference in location, as diverging trends have been seen within the same species when sampled in separate areas. An example is herring gull from the Baltic Sea for which PFOS increased during the entire study period 1991–2008, while in the same study a concentration peak was observed in 1994–2002 for herring gull from the North Sea (Rudel et al., 2011). Longer-chained PFCAs in bird eggs tend to increase globally in both marine, freshwater and terrestrial environments (Holmstrom et al., 2010; Ahrens et al., 2011; Braune and Letcher, 2013; Route et al., 2014). Most studies have focused on species in the marine environment. However, monitoring the freshwater and terrestrial environment provides a more accurate representation of recent exposure. In the marine food web, slow oceanic transport of persistent PFASs is a significant source of exposure, but integrates both historic and recent contamination (Prevedouros et al., 2006). Long-range atmospheric transport is of greater importance in the terrestrial and freshwater environments, since it is occurring at a faster rate than oceanic transport. Consequently, larger annual variations in the terrestrial environment have been observed compared to the marine environment (Bustnes et al., 2015). Volatile precursor compounds that are atmospherically transported can degrade to persistent PFASs (Ellis et al., 2004; Martin et al., 2006; Prevedouros et al., 2006). Common transformation intermediate products from several telomer precursors are fluorotelomer carboxylic acids (FTCAs) and fluorotelomer unsaturated carboxylic acids (FTUCAs). Another route of exposure in the freshwater and terrestrial environments is point-source emissions from waste-water treatment plants (WWTP), aqueous fire-fighting foam (AFFF) contamination, landfills, and manufacturing plants (Ahrens and Bundschuh, 2014). The influence of exposure from precursor compounds to the environment is not well studied. Differentiation between the food webs is important when assessing trends and patterns of PFASs. Stable isotope analysis of carbon and nitrogen (δ13C and δ15N) is a useful tool for elucidating dietary source and trophic level, thus illustrating how the food web is structured (Kelly, 2000). The δ15N value is associated with trophic level, where a δ15N shift of approximately 2–3‰ reflects one trophic level within a food web, while the δ13C value indicates feeding from the terrestrial/freshwater or marine environment. In this study, we investigated recent precursor exposure of PFASs and effects of production shifts by analyzing PFAS levels and assessing the composition profiles in bird eggs from the terrestrial and freshwater environments. Several PFAS groups were targeted; persistent PFASs that exist in their stable end form such as PFCAs and perfluorinated sulfonic acids (PFSAs), precursor compounds that are eventually subject to degradation (diPAPs, perfluorooctane sulfonamides (FOSAs), perfluorooctane sulfonamideoethanols (FOSEs), one fluorotelomer sulfonic acid (FTSA)), and intermediates (FTCAs, FTUCAs), formed during precursor transformation, together with stable isotopes of carbon and nitrogen. Eggs from three raptor species in terrestrial and freshwater environments were analyzed; tawny owl, common kestrel, and osprey. For the osprey samples a temporal trend was studied from 1997‐2001 to 2013.

2. Materials and methods 2.1. Samples Bird eggs (unhatchable) from three species were included in this study; osprey (Pandion haliaetus) common kestrel (Falco

41

tinnunculus), and tawny owl (Strix aluco). Osprey is a migratory bird and spend the winter season in West Africa, and feed mainly on fish (Glass and Watts, 2009). Common kestrel (Falco tinnunculus) is a migratory bird and spend the winter season in southwestern Europe during winter season, and feed on rodents (Costantini et al., 2005). Tawny owl is a non-migratory specie that feed mainly on rodents (Solonen and Karhunen, 2002). Ten osprey eggs from each period 1997–2001, 2007–2008, and 2013 were analyzed. A number of 40 eggs of common kestrel from 2014 were analyzed. Ten eggs of tawny owl from 2014 were analyzed. Samples were collected from different parts of Sweden in rural areas (Fig. S1, Table S1 and S2 SI). 2.2. Chemicals Standards of native PFCAs (C4-C14), 13C-labeled PFCAs (C4, C6, C8-C12), native PFSAs (C4, C6, C8, C10), technical PFOS, 13 18 C/ O-labeled PFSAs (C6, C8), native and 13C-labeled diPAPs (6:2, 8:2), native and 13C-labeled FTUCAs (6:2, 8:2, 10:2), native FTCAs (5:3, 7:3), native and labeled PFOSA, N-MeFOSA, N-EtFOSA, N-MeFOSE, N-EtFOSE, native and 13C-labeled 6:2 FTSA were obtained from Wellington Laboratories (Guelph, ON, Canada). Standard of native 10:2 diPAP was obtained from Chiron (Trondheim, Norway). HPLC-grade methanol and acetonitrile were purchased from Fisher Scientific (Ottawa, Canada), and Supelclean ENVI-carb (120/400 mesh) was purchased from Supelco (Bellafonte, PA, USA). Deionized water was from a purification system (18.2 Ω, Merck Millipore, Darmstadt, Germany) Ammonium hydroxide, glacial acetic acid, and sodium acetate were purchased from E. Merck (Darmstadt, Germany). Ammonium acetate was purchased from Fluka (Steinheim, Germany). 1-methyl piperidine was purchased from Sigma Aldrich (Stockholm, Sweden). 2.3. Extraction and clean-up An amount of 0.25 g thawed homogenate from one egg was added to a 15 mL polypropylene tube and spiked with labeled isotopic internal standards (2 ng PFCAs, PFSAs, FTUCAs, FTCAs, and FTSAs, 8 ng FOSAs, FOSEs, and diPAPs), followed by extraction with 4 mL acetonitrile. The sample was sonicated for 15 min, shaken for 15 min and centrifuged at 8000 g for 15 min. The supernatant was transferred to new polypropylene tubes with 50 mg ENVI-carb and 100 mL glacial acetic acid. Another volume of 4 mL acetonitrile was added to the sample and the extraction procedure was repeated and the aliquots were combined. After evaporation under nitrogen, the extracts were filtered (0.2 mm hydrophilic polypropylene filters) into LC-vials and the final volume was 200 mL. Labeled isotope performance standards were added and the extract was split into one fraction with final concentration of 80% acetonitrile and 20% Milli-Q water with 1-methylpiperidine (5 mM) and ammonium acetate (2 mM) for analysis of diPAPs and one fraction with 40% acetonitrile and 60% Milli-Q water with ammonium acetate (2 mM) for analysis of PFCAs, PFSAs, FTUCAs, FTCAs, FOSAs, FOSEs, and FTSAs. One procedural blank and quality control sample were included in each batch and treated in the same way as the samples. 2.4. Instrumental analysis For PFAS analysis two systems were used; one Acquity UPLC system coupled to a Quattro Premier XE mass spectrometer (Waters Corporation, Milford, USA), and one Acquity UPLC system coupled to a Xevo TQ-S mass spectrometer (Waters Corporation, Milford, USA). A guard column (PFC isolator, Waters Corporation, Milford, USA) was inserted between the pump and the injector to prevent contamination from the system. The analytes were

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separated using a 100 mm Acquity BEH C18 column (2.1 mm, 1.7 mm). Two methods with methanol and water mobile phases but with different additives were used. 1-methylpiperidine (5 mM) and ammonium acetate (2 mM) were used for diPAPs analysis while only 2 mM ammonium acetate was added for analysis of the remaining PFASs. The mass spectrometer was operated in negative electrospray with multiple reaction monitoring (MRM). Detailed description of instrumental settings can be found in Table S3, SI. Stable isotopes of nitrogen and carbon were analyzed using an elemental analyser (model EuroEA3024; Eurovector, Milan, Italy) coupled on line to an Isoprime isotope-ratio mass spectrometer (GV-Instruments, Manchester, UK). Results are expressed in the standard notation (δ13C and δ15N) in parts per thousand (‰) relative to their international standards Vienna Pee Dee Belemnite (V-PDB) and atmospheric N2, where δ13C or δ15N ¼((Rsample  Rstandard)/Rstandard)  1000 (‰), and R is the molar ratio 13C/12C or 15N/14N. 2.5. Quality control and quality assurance Quantification of PFASs was performed using isotope dilution. For compounds in which labeled internal standard was not available, the internal standard closest in retention time with similar analytical performance and from same compound class was used. For those diPAP homologues were no authentic standard were available, quantification was performed using standards of their structural isomers, and for quantification of x:2/x þ2:2 homologues, the closest diPAP homologue was used, and should be considered semi-quantified. Quantification of total PFOS was performed by summarization of linear PFOS and individual branched PFOS calculated against an external calibration curve using technical PFOS as standard. Recoveries were assessed using labeled internal standards, and a hen egg sample spiked with native compounds was included in each batch as a reference. The mean recoveries of labeled internal standards were in the range 68–95% for PFCAs, 91–93% for PFSAs, 56–89% for FTCAs and FTUCAs, 74% for PFOSA, 47–127% for FOSAs and FOSEs, and 135% for 6:2 FTSA. The recoveries of spiked native compounds in the hen egg were 87– 107% for PFCAs, 102–129% for PFSAs, 83–102% for FTCAs and FTUCAs, 38% for PFOSA, 85–118% for FOSAs and FOSEs, 105% for 6:2 FTSA, and 89–106% for diPAPs. For diPAPs the method limit of detection (MDL) was calculated as three times blank concentrations. For other PFASs the MDL was calculated as mean blank concentration with addition of three times the standard deviation. Fish muscle used in an interlaboratory study (ILS) was included as a quality control sample (n ¼6). The obtained concentrations for the control sample in this study were similar to the ILS reported values (z-scoreo1). Working standards consisting of wheat and bird eggs were used as reference material for stable isotope analysis, and standard deviation was less than 0.04 ‰ for δ13C and 0.19 ‰ for δ15N in 10 replicated samples.

2.6. Statistical analysis and calculations Statistical analysis was performed in Stata version 13. The PFAS levels were tested for skewness and kurtosis and were not found to be normally distributed (skewness 4 1.5, kurtosis 45.0). Therefore, nonparametric methods were chosen for statistical analyses. Spearman rank order correlations (rho) with Bonferroni adjusted significance level were used for analysis of correlations. Wilcoxon rank-sum test was performed to compare PFAS levels in different species and years.

3. Results and discussion 3.1. PFAS concentration PFOS was the predominating compound detected in all freshwater and terrestrial bird eggs, accounting for 72% of ∑PFAS in osprey, 45% in common kestrel, and 59% in tawny owl eggs (Fig. 1). The median PFOS level from 2013/2014 was significantly higher in osprey (70 ng/g) than in tawny owl (7.9 ng/g), which in turn was significantly higher than in common kestrel (3.8 ng/g) (Table 1). The contribution of linear PFOS (L-PFOS) to the total amount of PFOS was found to be 97% in osprey, 96% in tawny owl, and 88% in common kestrel. L-PFOS was highly enriched compared to commercial mixtures of PFOS produced with the electrochemical fluorination process (ECF), where L-PFOS contributes to around 70% (Benskin et al., 2010). Relatively high proportion of L-PFOS has previously been observed in birds and is probably a result of preferential accumulation of the linear isomer throughout the food web (Gebbink and Letcher, 2010). Other PFSAs than PFOS only had a small contribution to ∑PFSA. Perfluorohexane sulfonic acid (PFHxS) was found in a majority of the osprey eggs ( o0.16– 0.50 ng/g), while only four detections were observed in tawny owl (o0.16–0.96 ng/g) and two in common kestrel (o0.16–0.48 ng/g). Perfluorodecane sulfonic acid (PFDS) was detected in most of the ospreys (o 0.08–0.41 ng/g), and in three samples of tawny owl (0.12–0.25 ng/g). PFOS levels in osprey eggs were similar or lower than other freshwater species in previous studies, such as great cormorant (552 ng/g) and herring gull (292 ng/g) from Sweden and cormorant from Germany (400 ng/g) (Rudel et al., 2011; Norden et al., 2013). In the terrestrial environment, the PFOS levels in this study were in the same range as tawny owl eggs from Norway sampled in 2009 (Ahrens et al., 2011), and pigeon and rook eggs from Germany sampled in 2001–2008 (Rudel et al., 2011). Peregrine falcon eggs from Sweden sampled in 2006 had PFOS levels about ten times higher than terrestrial species in this study (Holmstrom et al., 2010), while PFOS was not detected in ptarmigan eggs from Greenland sampled in 2011 (Bossi et al., 2015). The longer-chained PFCAs, C9–C14, were the second most prevalent compound group and found in a majority of the eggs. For common kestrel, the ∑PFCAs exceeded the level of PFOS.

Fig. 1. PFAS profile in percentage of total PFAS in eggs from common kestrel, tawny owl, and osprey.

U. Eriksson et al. / Environmental Research 149 (2016) 40–47

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Table 1 Min, max and median concentration of PFCAs and PFSAs detected in 450% of common kestrel, tawny owl, and osprey eggs of (ng/g). Specie

Year

T-PFOS

L-PFOS

PFNA

PFDA

PFUnDA

PFDoDA

0.90 ( o 0.55–1.81)

1.4 ( o 0.15–3.2)

1.2 1.2 0.33 ( o 0.15–3.81) ( o 0.08–2.99) ( o 0.08–1.5)

Common kestrel (n ¼ 40) 2014

Median 3.8 Min-max (1.3–22.0)

3.4 (1.1–20.1)

0.52 ( o 0.12–1.2)

Tawny owl (n ¼10)

2014

Median 7.9 Min-max (3.5–34.9)

7.6 (3.2–33.6)

0.33 0.86 1.3 ( o 0.12–0.67) ( o 0.55–2.06) (0.53–2.4)

Osprey (n¼ 10)

2013

Median 69.7 Min-max (17.3–224)

67.6 (16.0–219)

1.5 (0.22–2.3)

7.5 (2.4–18.0)

Osprey (n¼ 10)

2008– 2009

Median 64.3 Min-max (12.8–337)

62.3 1.1 (12.4–333) (0.38–9.7)

Osprey (n¼ 10)

1997–2001

Median 103 Min-max (24.5– 667)

97 (23.7– 632)

0.88 (0.26–2.1)

Median ∑PFCA was 27.7 ng/g, 6.7 ng/g, and 4.6 ng/g for osprey, tawny owl, and common kestrel, respectively. For osprey and common kestrel, perfluoroundecanoic acid (PFUnDA) was the predominating PFCA (8.6% and 16% of ∑PFAS), while in tawny owl perfluorotridecanoic acid (PFTrDA) was predominating (11% of ∑PFAS). Long-chained PFCA median levels were about 5–10 times higher in osprey than in common kestrel and tawny owl. The predominance of long-chained PFCAs has previously been observed in for instance herring gulls from North America, and in cormorant and herring gulls from Sweden (Letcher et al., 2015; Norden et al., 2013). Of the other PFCAs, PFOA was detected in a small number of osprey ( o0.67–2.7 ng/g) and common kestrel eggs ( o0.67– 2.6 ng/g). The levels of the long-chained PFCAs seem to have increased in the terrestrial environment compared to the previous study from Norway on tawny owl eggs sampled in 2009 (Ahrens et al., 2011). The PFCA levels in osprey from 2013 were about five times lower than previous reported for freshwater avian species in Sweden (Norden et al., 2013). On the other hand, in cormorant from Germany, levels were similar to this study (Rudel et al., 2011). Variation in levels can be due to differences in diet between species, for instance, tawny owl and common kestrel feed mainly on rodents, while diet of peregrine falcon consist of medium-sized birds. Semi-persistent PFAS which environmental fate is degradation to persistent PFASs were detected at lower concentrations and less frequently than the persistent PFOS and PFCAs (Fig. 2). The intermediate transformation product 7:3 FTCA was found in all three species at levels ranging from o0.2–1.7 ng/g in osprey, o0.2– 1.5 ng/g in tawny owl, and o0.2–1.3 ng/g in common kestrel. Other fluorotelomer acids (5:3 FTCA, 8:2 FTUCA) were only detected in a two samples of osprey at levels of o0.02–0.06 ng/g and – o0.15–0.21 ng/g, respectively. Several diPAP homologues were detected in a limited number of osprey (n ¼3) and common kestrel (n ¼2) eggs (Table 2). PFOSA was only found in two samples of osprey, one from 1997 (1.5 ng/g) and one from 1998 (1.2 ng/g) Other FOSAs/FOSEs were not found in any sample. 6:2 FTSA was found in eggs from osprey and tawny owl, at levels ranging from o6.3–37 ng/g and o6.3–15 ng/g in samples from 2013/2014, respectively. 3.2. PFAS concentrations in relation to diet and trophic level While tawny owl and common kestrel feed on a strictly terrestrial diet, osprey can feed on fish from both the marine and

1.0 (0.48–2.40)

PFTrDA

PFTDA

1.4 (0.63–2.82)

0.55 ( o 0.08–2.5)

10.3 5.3 ( o 0.15–19.5) ( o 0.15–13.1)

4.6 ( o 0.08–7.5)

0.85 ( o 0.08–1.60)

7.6 (1.5–28.5)

11.3 (3.3–57.0)

3.7 (1.6–21.1)

5.6 (3.2–9.8)

0.55 (0.26–2.1)

3.2 (0.96–11.8)

5.1 (0.92–15.5)

2.0 (0.27–4.1)

1.7 (0.36–3.6)

0.15 ( o 0.08–0.38)

Fig. 2. Concentrations of semi-persistent PFASs in osprey, tawny owl, and common kestrel eggs (ng/g). For osprey eggs a temporal trend is given.

freshwater environment (Glass and Watts, 2009). Ospreys included in the present study fed on a freshwater diet rather than a marine, as indicated by the δ13C, ranging between  25.2‰ and  30.4‰ (Fig. 3). Birds with elements of marine diet have a δ13C 4 20‰ (Hobson, 1987). PFAS levels were higher in eggs from the freshwater environment, and within the terrestrial environment,

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Table 2 Minimum (ng/g), maximum (ng/g) and frequency of detection (%) of precursor and intermediate PFAS in common kestrel, tawny owl, and osprey eggs. Specie

Year

Common kestrel n ¼40

2014

Tawny owl n ¼ 10

Osprey n¼ 10

Osprey n¼ 10

Osprey n¼ 10

a

2014

2013

2008– 2009

1997– 2001

6:2 diPAP 6:2/8:2 diPAP

8:2 diPAP 6:2/10:2 diPAPa

8:2/10:2 diPAPa

10:2 diPAP 8:2/12:2 diPAPa

6:2 FTSA 7:3 FTCA

o6.3

Min o 0.04 Max 0.07 FOD (%) 5.1

o0.12 0.60 15

o0.04 0.16 5.1

o 0.03 0.19 7.7

o 0.05 0.51 13

o 0.08 0.84 10

o 0.01 1.4 5

Min o 0.02 Max FOD (%) 0

o0.08

o0.04

o 0.02

o 0.03

o 0.11

o 0.06

0

0

0

0

0

o0.01

o0.46 0.95 10

o 0.02 0.03 10

o 0.08 0.27 10

o 0.02 0.10 10 o 0.04 0.39 10

Min o 0.05 Max FOD (%) 0

0

Min o 0.05 Max FOD (%) 0

o0.01 0

o0.46 2.4 10

Min o 0.02 Max 0.39 FOD (%) 10

o0.01 0.45 10

o0.29 0.35 10

0

o 0.24 1.3 23

0

o6.3 15.3 10

o 0.24 1.5 20

o 0.90 1.4 10

o 0.98 1.3 10

o6.3 36.9 30

o 0.24 1.7 50

o 0.08 1.5 10

o 0.90 1.7 10

o 0.98 3.8 10

o6.3 52.2 80

o 0.24 2.7 80

o 0.17 0.28 10

o 0.52 1.1 10

o 1.09 1.2 10

o6.3 35.8 80

o 0.24 0.72 40

Semi-quantified concentrations.

the PFOS level was higher in tawny owl than common kestrel, which could be related to diet and trophic level. The trophic level was 2–3 timed higher for osprey than for common kestrel and tawny owl, revealed by the significantly higher δ15N in osprey (13.8‰) compared to tawny owl (7.4‰) and common kestrel (5.3‰) (Kelly, 2000). Tawny owl seem to feed on a higher trophic level than common kestrel, though this difference is not significant. The higher trophic level of osprey compared to common kestrel and tawny owl, though all three species are raptors, could be an effect of a more complex food web. The freshwater environment spans over more trophic levels than the terrestrial environment (Chase, 2000). Patterns of biomagnification have been shown to be similar in the freshwater and terrestrial environments with the highest bioaccumulation potential for PFCAs with chain length 9–11 (PFNA – PFUnDA) and PFOS, and the trophic magnification factors (TMF) of PFASs are similar in these environments (Muller et al., 2011; Xu et al., 2014). Accumulation in the aquatic environment due to the hydrophilic properties of PFAS may also contribute to observed differences between the osprey and the other birds. Within the terrestrial environment, the variation of δ15N was larger in common kestrel compared to tawny owl (variance ratio test, F39,9 ¼ 3.77, p o0.05), which indicate a broader choice of prey. This can be due to availability of voles and individual traits (Costantini et al., 2005). There’s a lack of studies of the effects on isotope signatures for addled bird eggs, but it has been assumed that stable isotope signatures are similar in fertilized and unfertilized bird eggs, since the formation of egg is not dependent upon fertilization (Hobson, 1995). Additionally, fresh and unhatchable eggs of marine turtles have been found to be isotopically equivalent (Ceriani et al., 2014). The use of unhatchable eggs rather than fresh eggs is more ethical and the only possible method to analyze vulnerable species. 3.3. PFAS profile The PFAS concentrations and homologue patterns differed between the freshwater and the terrestrial environment. The PFCA homologue pattern in the bird eggs differed between the two environmental compartments. The relative contribution of individual homologues to the ∑PFCA differed between species, with

Fig. 3. Mean values of stable isotopes signatures in eggs of common kestrel, tawny owl, and osprey. Error bars display 95% confidence interval.

longer-chained PFCAs (C12–C14) more abundant in the terrestrial avian species compared to osprey (Fig. S2, SI). For tawny owl and common kestrel, there was an odd-even chain-length pattern for the PFCAs, where PFUnDA 4PFDA, and PFTrDA 4PFDoDA, i.e. the even-numbered PFCA was higher than the adjacent shorter oddnumbered PFCA. In osprey the pattern differed, with almost equal contributions from PFDoDA and PFTrDA, and only slightly higher PFUnDA than PFDA. This odd-even pattern in the terrestrial environment could be attributed to significant input from atmospheric sources, were even-numbered fluorotelomer-based precursor compounds degrade to odd- and even-numbered PFCAs in similar yield, and increasing proportion of odd-numbered PFCAs is driven by bioaccumulation (Ellis et al., 2004; Armitage et al., 2009). Interestingly this odd-even pattern is less apparent in osprey eggs. Nevertheless, our findings are supported by similar results obtained in studies of other freshwater avian species. For instance, the median PFDA level was higher than PFUnDA in cormorant eggs from the Elbe estuary, sampled in 2009 (Rudel et al., 2011). Direct exposure from local sources may be more important for ospreys rather than indirect exposure from atmospheric degraded precursors. Direct exposure of precursor compounds to osprey and common kestrel was also revealed by the prevalence of

U. Eriksson et al. / Environmental Research 149 (2016) 40–47

diPAPs in these species. DiPAPs are non-volatile compounds presumably unstable in biological organisms and their presence in bird eggs are therefore likely attributed to local sources in the terrestrial and freshwater environment, though it’s possible that they can be long-range transported atmospherically on particles. In waste-water treatment plant sludge, diPAP levels of up to 2100 ng/g have been detected, and the application of sludge in agriculture can further spread these compounds into the environment (D'eon et al., 2009; Lee et al., 2014). The levels of diPAPs were higher in the freshwater environment than the terrestrial environment. The highest level of 6:2 diPAP was found in one osprey sample from 1999 (0.39 ng/g), and the highest level of 8:2 diPAP was found in one osprey sample from 2008 (2.4 ng/g). DiPAPs have shown to biodegrade into PFCAs in rats, which could explain the low levels and frequency of detection in the present study (D'eon and Mabury, 2007). While 6:2 FTSA was detected in only one sample from the terrestrial environment, it was frequently occurring in the freshwater environment and detected in 63% of the osprey eggs. The prevalent occurrence of 6:2 FTSA could be a result of local AFFF contamination and subsequent bioaccumulation. FTSAs have been suggested to be degradation products of ingredients in AFFF, such as fluoroalkylthioamide sulfonates (FTSASs), and can further degrade into PFCAs (Wang et al., 2011; Houtz et al., 2013). The levels of 6:2 FTSA indicate that bioaccumulation occurs in the osprey. Although it has been suggested that 6:2 FTSA is rapidly eliminated and biotransformed in fish, it has shown to bioaccumulate in midge larvae (Yeung and Mabury, 2013; Bertin et al., 2014). There are few reports of the less studied semi-persistent PFASs in wildlife. However, 7:3 FTCA has been found in birds from freshwater and terrestrial environments in Japan (Guruge et al., 2011). Liver samples were analyzed and ranged from 0.25 to 62 ng/ g, which were generally higher than in this study. However, it should be noted that PFAS profiles are tissue-specific with large variation between egg and liver (Norden et al., 2013), therefore these results are not directly comparable. 6:2 diPAP and 8:2 diPAP have been found in lake trout from Great Lakes in Canada, at levels ranging from non-detectable (n.d.) to 0.4 ng/g (Guo et al., 2012). Recently, 8:2 diPAP was found in tuna from the Indian Ocean at levels of 11.4–11.5 ng/g (Zabaleta et al., 2015). 6:2 FTSA has been detected in fish, benthic worms and benthic invertebrates (Kärrman et al., 2011; Loi et al., 2013; Lescord et al., 2015). Mean levels of 6:2 FTSA has been reported to be n.d. – 0.43 ng/g in benthic invertebrates and n.d. – 0.97 ng/g in juvenile char from the Canadian Arctic (Lescord et al., 2015). PFOSA has shown to decrease from 1999 to 2008 in gulls from Canada, which is in agreement with the present study (Gebbink et al., 2011a; Letcher et al., 2015). 3.4. Temporal trend in osprey eggs No significant difference was seen in PFOS levels in osprey eggs from 1997‐2001 (103 ng/g), 2008‐2009 (64 ng/g), and 2013 (70 ng/ g) (Fig. 4). A decline was expected as a result of the phase-out and regulatory interventions, however, continued high levels may be related to the persistence and long half live of PFOS, emissions from historically contaminated sites, and/or increased and ongoing production in Asia (Wang et al., 2014b). Other studies show inconclusive results for PFOS concentrations. The present study showed a similar trend for PFOS as observed in peregrine falcon eggs from Sweden, where no change in concentration was observed between 1984 and 2006 (Holmstrom et al., 2010). No significant trend of PFOS levels was observed between 1975 and 2011 in thick-billed murre and northern fulmar eggs from the marine arctic environment (Braune and Letcher, 2013). Other studies have shown decreasing levels of PFOS, which

45

PFOS 800 600 400 200

1997-2001

2008-2009

2013

Fig. 4. Temporal trend of PFOS in osprey eggs. Boxes display median, 2nd quartile, and 3rd quartile. Whiskers display minimum and maximum values, outliers defined as 4 1.5 interquartile range excluded. Y-axis is in logarithmic scale.

was found in for example tawny owl eggs from Norway between 1986 and 2009, and herring gull eggs from Canada between 1990 and 2012/2013 (Ahrens et al., 2011; Gebbink et al., 2011b; Letcher et al., 2015). Species-specific trends were reported in a Canadian study, with peaking PFOS levels around 2000 for rhinoceros auklet and double crested cormorant eggs, while no change was observed in great blue heron eggs, and increasing concentrations in stormpetrel egg between 1990 and 2011 (Miller et al., 2015). There was a significant increase for PFDA, PFUnDA, PFDoDA, PFTrDA, and PFTDA between 1997‐2001 and 2008‐2009 (Fig. 5). Thereafter, between 2008‐2009 and 2013, there was no significant change in PFCA concentrations. Over the years 1997–2013 the proportional amount of PFDA increased in osprey, while the proportion of PFUnDA decreased. The trend of increasing long-chained PFCA levels has previously been observed for other birds for example in eggs of tawny owl (Ahrens et al., 2011), and peregrine falcons (Holmstrom et al., 2010). More recently, PFTrDA and PFTDA have increased in bald eagles from US between 2008 and 2011 (Route et al., 2014). An increase of long-chained PFCAs between 2000 and 2006 could be related to the production shift from the POSF-based production to the telomerization process (Wang et al., 2014a) However, in 2006, the PFOA stewardship program was launched, aiming to completely eliminate long-chained PFCAs and their precursors in 2015 (EPA 2006). Therefore, the levels of PFCAs are expected to decrease, yet this effect is not observed in the ospreys. No temporal trend could be determined for the semipersistant PFASs detected in osprey eggs, largely hampered by the low frequency of detection. It's notable though that the diPAPs, 6:2 FTSA and 7:3 FTCA were detectable throughout the whole study period. Less frequent detection of 6:2 FTSA in 2013 compared to previous years might indicate a decrement. Variations in 6:2 FTSA can also be a result of different sampling locations between years, as 6:2 FTSA levels are likely affected by local point sources. Further studies are needed to evaluate the trends of the semipersistant PFAS. 3.5. Geographical variations The levels of PFAS were generally higher in the south and the west parts of Sweden (Table S4, SI). In the south, tawny owl had significantly higher levels of PFOS (p o0.05) compared to the west. Common kestrel also had significantly higher levels of PFOS (p o0.05) in the south compared to the central and northern parts of the country. For tawny owl there was a tendency of higher levels of ∑PFCA in the south and west part compared to the south-east

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U. Eriksson et al. / Environmental Research 149 (2016) 40–47 PFNA

10 8 6 4

PFDA

PFUnDA

30

60 40

20

20

10

2

1997 - 2001

2008 - 2009

2013

PFDoDA

1997 - 2001

2008 - 2009

2008 - 2009

2013

PFTDA

2 1.5 1

4

5

1997 - 2001

PFTrDA

10 8 6

20 15 10

2013

.5

2

1997 - 2001

2008 - 2009

2013

1997 -1999

2008 - 2009

2013

1997 - 2001

2008 - 2009

2013

Fig. 5. Temporal trend of PFCAs in osprey eggs (ng/g, n ¼10 for each time point). Boxes display median, 2nd quartile, and 3rd quartile. Whiskers display minimum and maximum values, outliers defined as 41.5 interquartile range excluded. Y-axis is in logarithmic scale.

part, and for common kestrel the ∑PFCA levels were higher in the southern parts than the north part. Interestingly, the δ15N in tawny owl was significantly higher in the south and west of Sweden, and for common kestrel, the δ15N was significantly higher in the south part and significantly lower in the north part of Sweden. Baseline values are not known, and therefore geographical differences of δ15N should be interpreted with caution. There are several possible reasons for higher PFAS levels and δ15N in the south of Sweden. The higher δ15N in bird eggs could be due to different diet habits, in terms of feeding on a higher trophic level thus leading to increased PFOS levels. Another explanation for higher PFAS levels could be higher amount of precipitation in the south, leading to increased atmospheric input of PFASs. Environmental conditions such as precipitation have shown to be positively associated with PFAS levels in tawny owl from Norway (Bustnes et al., 2015). Higher δ15N in the south can be caused by nitrogen input through intensive agriculture practices in these areas compared to the northern Sweden, which consequently can explain the positive relationship between δ15N and PFAS levels. Other reasons for higher PFOS levels in the southern part of Sweden could be from more frequent point-source emissions, for instance firefighting practice grounds. 3.6. Implications The levels of PFOS in the osprey eggs are in the range where adverse effects such as reduced hatchability have been observed. A value of 100 ng PFOS/g egg is reported as lowest-observable-adverse-effect (LOAEL) for PFOS in white leghorn chicken, based on reduced hatchability (Molina et al., 2006). In this study, 40% of the osprey eggs from 2013 exceed this level. If hatchability is affected by the PFAS levels, it cannot be ruled out that the use of unhatchable eggs in this study may pose a positive bias to the results, Other toxicological effects of PFOS observed in laboratory bird studies includes pathological changes in the liver (Molina et al., 2006), immunological, morphological, and neurological effects (Peden-Adams et al., 2009), and increased fatty acid oxidation (Norden et al., 2012). The toxicological effects of long-chained PFCAs in birds are much less studied than PFOS and PFOA. However, several negative effects have been linked to long-chained PFCA exposure. Some examples include the relationship between PFNA and male body-

condition, lower hatching success related to PFDoDA, and decreased glucocorticoid hormones related to PFTrDA and PFTDA, observed in black-legged kittiwakes (Tartu et al., 2014). For semipersistent PFASs, there’s even less information available about toxicological effects, and to our knowledge, no studies on birds have been published. It has been shown that diPAPs affect steroidogenesis and alter androgen and estrogen levels (Rosenmai et al., 2013). 6:2 FTSA has been found to be more toxic than PFOA in rainbow trout, and have similar toxicity to PFOS in green algae (Hoke et al., 2015). FTCAs and FTUCAs have shown to be several times more toxic than PFCAs in acute and chronic tests with freshwater invertebrates (Phillips et al., 2007). PFOS alone poses a risk to birds in the freshwater environment and the response to efforts reducing the emissions is slow as seen in the present study. Continuous high and increasing levels of PFCAs, prevalence of precursor compounds, and the unknown amount of other emerging compounds, represents additional hazards that are of great concern. As has been declared in the Madrid Statement, signed by hundreds of scientists, the use of PFAS and their fluorinated alternatives needs to be reduced, in order to avoid long-term harm to human health and the environment (Blum et al., 2015). Further studies on contaminants compositions and their levels are needed in order to understand the effect of regulations of PFAS and use of their replacement compounds.

Acknowledgments This project was funded by the Swedish Environmental Protection Agency (SEPA) (Grant number 2220-14-004). We would like to acknowledge Henrik Dahlgren, Douglas Jones, Eva Kylberg, Malin Stenström, and Lina Jansson at the Swedish Museum of Natural History (SMNH), and all the bird banners that participated in the study. We would also like to acknowledge Jan Sondell and Tjelvar Odsjö for their long term study of ospreys in Sweden, including collecting dead eggs. The project started already in 1970. We also want to thank Katarina Loso at SMNH for help with the 2011 survey, and Jordan Stubleski at Örebro University for proof reading.

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Appendix A. Supplementary material Supplementary data associated with this article can be found in the online version at http://dx.doi.org/10.1016/j.envres.2016.04. 038.

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