Perfluoroalkyl substances (PFASs) in the marine environment: Spatial distribution and temporal profile shifts in shellfish from French coasts

Chemosphere 228 (2019) 640e648

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Perfluoroalkyl substances (PFASs) in the marine environment: Spatial distribution and temporal profile shifts in shellfish from French coasts
ge, Pollono Charles, Aminot Yann Munschy Catherine*, Bely Nade IFREMER (Institut Français de Recherche pour l’Exploitation de la Mer), Laboratory of Biogeochemistry of Organic Contaminants, Rue de l’Ile d’Yeu, BP 21105, Nantes Cedex 3, 44311, France

h i g h l i g h t s
PFAS levels and profiles investigated in shellfish from French coasts in 2013e2017.
Evidence of current widespread contamination by PFOS and long-chain PFCAs. P
Higher levels of PFOS in main estuaries and of PFCA at one industrial site.
Profile shifts from PFOS to long-chain PFCAs were evidenced over the study period.
Significant decrease in PFOS concentrations after mid-90s.

a r t i c l e i n f o

a b s t r a c t

Article history: Received 24 February 2019 Received in revised form 26 April 2019 Accepted 27 April 2019 Available online 29 April 2019

Perfluoroalkyl substances (PFASs) were investigated in filter-feeding shellfish collected from 2013 to 2017 along the English Channel, Atlantic and Mediterranean coasts of France. PFOS (perfluorooctane sulfonate), PFTrDA (perfluorotridecanoic acid), PFTeDA (perfluorotetradecanoic acid), PFDoDA (perfluorododecanoic acid) and PFUnDA (perfluoroundecanoic acid) were detected in more than 80% of samples, thus indicating widespread contamination of the French coastal environment by these chemicals. The distribution of PFAS concentrations showed differences according to sampling locations and years. PFOS was the predominant PFAS in most samples collected from English Channel and Atlantic coasts until 2014, but the opposite was observed in 2015, 2016 and 2017, while perfluoroalkyl carboxylic acids (PFCAs) prevailed in Mediterranean samples in all study years. Among PFCAs, PFTrDA showed the highest maximum (1.36 ng g1 ww) and median (0.077 ng g1 ww) concentrations in 2016e2017. Other PFAS median concentrations were within the 0.014 (PFNA) - 0.055 (PFTeDA) ng g1 ww range. The profiles determined each year in most Mediterranean samples suggest distinctive sources. PFOS median concentrations showed a significant decrease over the study years, from 0.118 to 0.126 ng g1 ww in 2013 P e2015 to 0.066 ng g1 ww in 2016 and 2017. PFCAs showed no trends in concentration ranges over the same years. The shift in PFAS profiles from PFOS to long-chain PFCAs over the study period reflects PFOS production phase-out, combined with continuous inputs of PFCAs into the marine environment. These results provide reference data for future studies of the occurrence of contaminants of emerging concern on European coasts. © 2019 Elsevier Ltd. All rights reserved.

Handling Editor: Prof. J. de Boer Keywords: PFOS PFCAs Shellfish Levels Profiles Temporal trends

1. Introduction Perfluoroalkyl substances (PFASs) refer to a major class of chemical products, including the more frequently-studied perfluoroalkane sulfonic acids (PFSAs, including perfluorooctane

* Corresponding author. E-mail address: [email protected] (M. Catherine). https://doi.org/10.1016/j.chemosphere.2019.04.205 0045-6535/© 2019 Elsevier Ltd. All rights reserved.

sulfonate ePFOS) and perfluoroalkyl carboxylic acids (PFCAs, which include perfluorooctanoic acid ePFOA). Due to their hydrophobic and lipophobic properties, PFASs have been widely used as surfactants, lubricants, coating agents, stain repellents, dispersants and polishes in a vast number of industrial applications, consumer products and fire-fighting foams (Prevedouros et al., 2006). PFAS inputs into the environment occur through direct release (from manufacture, use and disposal of products containing PFASs or via their unintentional occurrence as impurities in industrial and

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domestic activities), or indirect release (through abiotic or biotic degradation of their precursors) (Armitage et al., 2006; Paul et al., 2009; Buck et al., 2011; Wang et al., 2014a, 2014b). PFASs have been produced since the 1950s; after 2002, their global production tended to shift from long-chain PFASs to shorter-chain homologues or other non-fluorinated substances with expected lower toxicities (Buhrke et al., 2013; Wang et al. 2013, 2015). Indeed, long-chain PFASs (i.e., PFCAs with 8 carbons and more, PFSAs with 6 carbons and more) represent a threat for living organisms as they are more bioaccumulative than their short-chain counterparts and biomagnify in the food chain (Conder et al., 2008; Kelly et al., 2009; Munoz et al., 2017). PFASs share similar properties to persistent organic pollutants (POPs), i.e., persistence, bioaccumulation in organisms, toxicity and global distribution in the environment (Giesy and Kannan, 2001, 2002; Lau et al., 2007; Ahrens, 2011; Houde et al., 2011). As a result, PFOS and its related compounds were included in the Stockholm Convention list of POPs in 2009 (UNEP, 2017). In Europe, the use of PFOS and its derivatives has been restricted since 2007 (EU, 2006); the European Union Water Framework Directive added them to its priority substances watchlist in 2013 (European Parliament, 2013). Moreover, C11eC14 PFCAs were identified as being highly-persistent and bioaccumulative and added to the candidate list of Substances of Very High Concern for authorisation under the European Registration, Evaluation, Authorisation and Restriction of Chemicals (REACH) regulation (EG1907/2006) (ECHA (European Chemical Agency), 2013). Despite regulations on PFOS and PFCAs, many PFASs are still released into the environment, especially in Asia (Wang et al., 2014a, 2014b; Su et al., 2018). Oceanic currents and atmospheric circulation are the main drivers behind the global transportation of directly or indirectlyemitted PFAS compounds (Armitage et al., 2006). On a more local scale, waste water treatment plants (WWTPs) and industrial effluents are reported to be the main sources of PFASs in the aquatic environment (Sinclair and Kannan, 2006; Bossi et al., 2008; Clara et al., 2008; S anchez-Avila et al., 2010). These discharges eventually end up in the marine environment via rivers and runoffs; as a result, oceanic waters have become the main PFAS reservoir (Yamashita et al., 2008). The contamination of marine ecosystems by PFASs is therefore of high concern: PFASs are recognized as

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exerting various toxic effects on fish, such as endocrine disruption and hepatotoxicity (Hoff et al., 2003; Liu et al., 2007; Wei et al., 2007; Mortensen et al., 2011). Moreover, seafood has been reported as the primary dietary source of human exposure to PFASs (Falandysz et al., 2006; Ericson et al., 2008; Haug et al., 2010; Domingo et al., 2012). This study aimed to investigate the contamination levels and profiles of PFASs and their spatial distribution along French coastlines using marine shellfish as sentinel species of coastal environment chemical contamination. Referred to as the “Mussel Watch Program”, this monitoring strategy, together with environmental specimen banks, have been widely used for many years in various countries worldwide for legacy contaminants (O'Connor and Lauenstein, 2006; Ramu et al., 2007; Tanabe et al., 2007) and more recently extended to various contaminants of emerging concern. Shellfish have been shown to be extremely good bioindicators of PFAS contamination (Munschy et al., 2013, 2015; Maruya et al., 2014). This study provides new results on the contamination of the French coastal marine environment by PFASs 15 years after the first PFOS restrictions and provides a valuable comparison with previously-obtained data (Munschy et al., 2013, 2015). More broadly, it also provides an update on the levels and temporal trends of contaminants of emerging concern in European coastal waters. 2. Materials and methods 2.1. Sampling strategy The mussel (Mytilus edulis, Mytilus galloprovincialis) and oyster (Crassostrea gigas) samples used for our study were collected once a year from 2013 to 2017 at selected locations in the English Channel, Atlantic and along Mediterranean coasts (except in 2015), as shown in Fig. 1. These locations cover main estuaries and deltas (Seine, ^ne), together with smaller tributaries. In addiLoire, Gironde, Rho tion, samples obtained from IFREMER's Environmental Specimen Bank were used to study temporal trends at selected sites. The shellfish were collected and handled in accordance with international guidelines for the monitoring of contaminants in

Fig. 1. Sampling sites for study shellfish collected in 2016e2017 from the English Channel, Atlantic and Mediterranean coasts.

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biota (OSPAR, 2009), as previously described (Munschy et al., 2013, 2015). Briefly, all samples were collected in the same manner and at the same time of year (from late November to early December); each individual had spent at least 6 months on site beforehand. Each composite sample consisted of at least 50 ± 5 mussels of similar size (30e60 mm shell length) or 10 ± 1 oysters (90e140 mm shell length). All samples were systematically depurated in filtered water for 24 h, shelled, homogenized (whole soft body) and stored at 20  C prior to freeze-drying. The dried samples were stored in a cool, dry, dark place until further analysis. 2.2. Chemicals Methanol (MeOH) was Picograde® quality and provided by LGC Promochem (Wesel, Germany). Reagents, including ammonium acetate (CH3COONH4, 98.0%), ammonium hydroxide (NH4OH, 32%), potassium hydroxide (KOH,  85%) and glacial acetic acid (CH3COOH,  99.8%), were supplied by Merck (Darmstadt, Germany). Ultrapure water (>14 MU cm) was produced using an € ttingen, arium® mini water purification system from Sartorius (Go Germany). High purity argon gas (>99.999%) was used as collision gas. Desolvation and nebulizing gas were supplied by a nitrogen generator NM32LA from Peak Scientific (Inchinnan, Scotland, UK). SPE WAX cartridges (150 mg, 6 mL) were purchased from Waters Corp. (Milford, MA, USA) and Envicarb cartridges (500 mg, 6 mL) from Supelco (Sigma-Aldrich, Saint-Quentin Fallavier, France). PFAS and corresponding labelled internal standards were obtained from Wellington Laboratories (Guelph, Canada). The following compounds, including five C4- to C10-perfluoroalkyl sulfonates (PFSAs) and nine C6- to C14 perfluorocarboxylic acids (PFCAs), were analysed: perfluorobutane sulfonate (PFBS); perfluorohexane sulfonate (PFHxS); perfluoroheptane sulfonate (PFHpS); perfluorooctane sulfonate (PFOS); perfluorodecane sulfonate (PFDS); perfluorohexanoic acid (PFHxA); perfluoroheptanoic acid (PFHpA); perfluorooctanoic acid (PFOA); perfluorononanoic acid (PFNA); perfluorodecanoic acid (PFDA); perfluoroundecanoic acid (PFUnDA); perfluorododecanoic acid (PFDoDA); perfluorotridecanoic acid (PFTrDA) and perfluorotetradecanoic acid (PFTeDA). Labelled compounds used as internal standards (surrogates) for quantification by isotopic dilution were added before extraction to each sample, blank and internal QA/QC material. Targeted analytes were quantified using their corresponding isotope labelled standard, unless otherwise stated. The labelled standards were PFHxS 18 O2 (used to quantify PFBS and PFHxS), PFOS 13C4 (used to quantify PFHpS, PFOS and PFDS), PFHxA 13C2 (used to quantify PFHxA and PFHpA), PFOA 13C4, PFNA 13C5, PFDA 13C2, PFUnDA 13C2, PFDoDA 13 C2, and PFTeDA 13C2 (used to quantify PFTrDA and PFTeDA). PFOS 13 C8 was added to the purified extracts before injection and used as an injection standard. 2.3. Sample preparation and analysis One gram of a freeze-dried sample was transferred to a 15 mL polypropylene tube and an internal standard mixture of nine labelled compounds was added prior to agitation. A liquid solid extraction (LSE) was performed using 15 mL of a blend of MeOH/ KOH (0.01 M of KOH). The sample was then mechanically agitated and left in contact for one night. The supernatant was collected and purified onto two consecutive SPE cartridges: a WAX weak anion exchange stationary phase (Waters® Oasis Wax, 150 mg, 6 mL) eluted with MeOH/NH4OH (99.5:0.5, v/v) and an Envicarb charcoal stationary phase (Supelco® Envi Carb, 500 mg, 6 mL), eluted with MeOH:acetic acid (80:1, v/v) (Couderc et al., 2015). Lastly, the

extracts were evaporated to dryness under a gentle stream of nitrogen at ambient temperature, then reconstituted in 200 mL of a mixture of MeOH:H2O (50:50, v/v) to which PFOS 13C8 was added. Analysis was performed using an Acquity ultra performance liquid chromatograph (UPLC®, Waters Corp.) coupled to a triple quadrupole mass spectrometer (Xevo® TQ-S micro, Waters Corp.) interfaced with an electrospray ionisation source Z-spray™ (Waters Corp.); UPLC separation was achieved using an Acquity UPLC BEH C18 reversed-phase column (1.7 mm, 50  2.1 mm, Waters Corp.). Elution solvents were ammonium acetate in water (20 mM) (A) and methanol (B). The mobile phase flow rate was 0.5 mL min1. The gradient started at 25% B for 0.5 min, was increased to 85% in 4.5 min and 100% in 0.1 min, held for 0.9 min, returned to initial conditions in 0.1 min and held for 2.9 min. A volume of 10 mL of sample was injected with an automatic injector. The mass spectrometer was operated in negative ionisation mode using multiple reaction monitoring with argon as the collision gas. Two transitions were recorded per analyte. The capillary voltage was 2.8 kV. The source temperature and probe temperature were 150  C and 500  C, respectively. Nitrogen was used as the nebulizing and desolvation gas. Instrumental operations, data acquisition and peak integration were performed with MassLynx Software (v.4.1, Waters). 2.4. QA/QC PFASs were quantified by isotopic dilution using 13C-labelled compounds and PFHxS 18O2, used to quantify PFBS and PFHxS. A six-point calibration curve ranging from 0.05 pg mL1 to 100 pg mL1 was used to calculate relative response factors and check linearity. Laboratory blanks were simultaneously processed and monitored in parallel with the samples, and the signal of each compound in the blanks was checked to assess contamination throughout the analytical procedure. A commercially purchased mussel sample sequentially shelled, homogenized, freeze-dried and spiked at 0.2 ng g1 dw of each target PFAS was used as an in-house quality control (QC) and included in each series of analyses to assure repeatability. The results obtained for this QC were used to set up a quality control chart guaranteeing the robustness of the entire analytical procedure. In addition, the laboratory regularly takes part in the QUASIMEME (Quality Assurance of Information for Marine Environmental Monitoring in Europe) intercomparison exercises for PFAS, and obtained satisfactory Z scores, i.e., between 1.1 and þ 1.3, on a 2018 mussel sample. For the 2017 samples, labelled standard recoveries were above 60% (between 65 ± 11% for 13C-PFHxA and 100 ± 11% for 13C-PFOS), except for 13CPFTeDA (43 ± 23%). Laboratory blank levels were between 0.004 ng g1 ww (PFHpA) and 0.024 ng g1 ww (PFOA). LOQ values were determined for each target compound in each analysed sample according to Wenzl et al. (2016), i.e., using a signal-to-noise ratio of 3 (peak-to-peak) for the less intensive raw data signal(qualifier ion)). Mean LOQs were between 0.005 ng g1 ww (PFHpA) and 0.042 ng g1 ww (PFBS). The relative standard deviations of the in-house quality control mussel versus target values were between 5% (PFDA) and 27% (PFDS). Extended uncertainties derived from the method validation were between 24% (PFHpS) and 50% (PFTrDA). Concentrations in samples were blank-corrected and expressed on P a wet weight basis. PFCA concentrations refer to the sum of compounds from PFNA to PFTeDA, with concentrations below LOQs counted as zero. 2.5. Statistical analysis Statistical analyses were performed using StatSoft Statistica software v7.1. Correlations were tested using simple linear

M. Catherine et al. / Chemosphere 228 (2019) 640e648

regression coefficients. In view of the limited number of samples per group, data comparisons (concentrations and ratios) were performed using non-parametric tests (ManneWhitney test for comparison of two independent samples or one-way ANOVA Kruskal-Wallis's test) with a significance level of 0.05. Results were considered to be significant only when both tests were significant. The significance of the temporal trend was assessed using the nonparametric Mann-Kendall statistical method, which assesses ranks across samples. 3. Results and discussion 3.1. Current (2016e2017) PFAS levels PFOS was the only PFSA determined above LOQs in samples collected in 2016 and 2017, with a detection frequency of 100% and concentrations in the 0.007e0.218 ng g1 ww range (median value of 0.066 ng g1 ww). PFOS is reportedly the main PFAS found in all species of aquatic biota throughout the world (Martin et al., 2004; Houde et al., 2011), although some studies have reported PFOA as the predominant compound, in particular in shellfish from Asia, in relation to direct sources (Nakata et al., 2006; Pan et al., 2010). However, in our samples, PFOA was not detected at all, revealing an absence of significant direct inputs on a national scale. In addition, this result is consistent with PFOA's lower bioaccumulation propensity versus PFOS or long-chain PFCAs (Martin et al., 2003). In 2016e2017, PFOS mean concentrations above the nationwide median were found at 3 sites in the English Channel (Somme Bay, Antifer, Seine estuary), seven sites on the Atlantic coast (Vilaine , Loire estuary, Bourgneuf Bay, Charente estuary, estuary, Pen Be Gironde estuary) and at Thau lagoon in the anean Sea (Fig. 2). Interestingly, the previously-identified maximum PFOS concentrations found on coastlines, reported at the Loire estuary site in samples collected in 2010 and 2011 (0.87 ng g1 ww and 0.68 ng g1 ww, respectively, Munschy et al., 2013), were 4e7 times lower in 2016 and 2017 (0.119 ng g1 ww and 0.161 ng g1 ww, respectively), although PFOS concentrations at this site remain among the highest across all samples. PFOS environmental emissions mainly result from diffuse sources via direct release from

Fig. 2. PFOS (left) and

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consumer products (Paul et al., 2009) and its discharge into rivers has been related to population levels (Pistocchi and Loos, 2009). However, according to our data set, PFOS concentrations did not always correlate with population, e.g., concentrations found in the Bay of Marseille were among the lowest found along coasts (Fig. 2), despite Marseille being France's 2nd most populated city, with over 850,000 inhabitants. PFOS has also been found in effluents from the metal and paper industries (Clara et al., 2008). Among PFCAs, longer-chain PFTrDA and PFTeDA were identified in 100% of samples at concentrations in the 0.018e1.364 ng g1 ww and 0.014e0.667 ng g1 ww ranges, respectively. Other PFCAs, namely PFDoDA, PFDA, PFUnDA and PFNA, were detected in 88%, 83%, 78% and 58% of the samples, at median concentrations of 0.025 ng g1 (0.007e0.259 ng g1 ww range), 0.016 ng g1 (0.003e0.046 ng g1 range), 0.017 ng g1 (0.006e0.105 ng g1 range) and 0.014 ng g1 (0.008e0.068 ng g1 range), respectively. These results contrast with our previously-reported data, which pinpointed the detection of PFCAs mainly in Mediterranean samples and, among them, the occurrence of PFUnDA in Mediterranean samples only (Munschy et al., 2013, 2015). PFCA mean concentrations (2016e2017) above the median were determined at Antifer, , in the Seine estuary (EC), in the Loire, Gironde and Nivelle Pen Be estuaries (Atlantic) and at all sites except Corsica on the Mediterranean coast (Fig. 2). P PFCA maximum concentrations were determined at the Gulf of Fos site on the Mediterranean coast in both 2016 and 2017 (1.46 ng g1 ww and 2.37 ng g1 ww, respectively). The Gulf of Fos is a semi-enclosed bay bordered by a heavily-industrialized zone, home to chemical, petroleum and steel-work plants (Mille et al., 2007). Hence, the PFCAs identified in mussel samples from this site are very likely to be of local industrial origin, although the ^ ne river with particular mixed influence of inputs from the Rho long-chain PFAS profiles, already reported in fish and sediment
ge et al., 2012), cannot be ruled out. PFCAs (Bertin et al., 2014; Mie have previously been identified at high concentrations in effluent waters from the textile industry, but they may also originate from other industrial activities (Bossi et al., 2008; Clara et al., 2008). Their sources are also reported to be indirect, via the degradation of precursors such as fluorotelomer alcohols (FTOHs) (Ellis et al.,

P PFCA (right) concentrations (ng g1 ww) in shellfish collected in 2016 (black bars) and 2017 (grey bars) along French coasts. Est. ¼ estuary.

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2004), or resulting from the manufacture of fluorinated polymers (Martin et al., 2004). In our samples, PFTrDA and PFTeDA showed the highest concentrations of all PFCAs (3e6 times higher than other PFCAs); this is consistent with the increased bioaccumulation propensity of longer carbon chain length PFASs. These two compounds were highly correlated (r ¼ 0.93, p < 0.05), suggesting similar sources and behaviours in shellfish. PFTrDA in sediment has previously been attributed to fluoropolymer and polyvinylidene fluoride manufacturing plants, along with PFUnDA (Dauchy et al., 2012). However in our samples, PFUnDA did not correlate with PFTrDA. PFTrDA and PFTeDA also correlated significantly with PFDoDA (r ¼ 0.86 and 0.95, respectively, p < 0.05), which has been reported as originating from stain and greaseproof coatings on various consumer products in the US (Maruya et al., 2014). PFCAs with odd carbon numbers were found in higher proportions than corresponding shorter, even-chain length PFCAs (e.g., PFUnDA > PFDA, PFTrDA > PFDoDA) at most sites. The predominance of odd-chain length PFCAs has previously been observed in various marine species including fish, seabirds and mammals (Sturm and Ahrens, 2010; Pan et al., 2010) and related to

atmospheric sources via the degradation of FTOHs, together with higher bioaccumulation propensities (Martin et al., 2004).

3.2. Comparison with worldwide levels Data on PFOS and PFCA concentrations in filter-feeding shellfish from various locations is presented in Table 1. In view of data availability in the literature, PFOS data was restricted to Europe, while PFCA data was extended worldwide. To our knowledge, very few studies have reported PFAS presence in shellfish in the French marine environment (Munschy et al., 2013, 2015; Munoz et al., 2017), while PFOS has been reported below LODs or at low levels in shellfish from various European countries (Bossi et al., 2008; mez et al., Nania et al., 2009; Fern andez-Sanjuan et al., 2010; Go 2011). This could be assignable to a PFOS short depuration halflife (87e131 h), as recently shown in aquatic invertebrates such as oysters (O'Connor et al., 2018) and crabs (Taylor et al., 2017). With this in mind, our results strongly suggest that the PFAS contamination observed in shellfish on French coasts results from continuous inputs rather than point sources.

Table 1 PFOS and PFCA concentration ranges (minimum-maximum/median or mean) in shellfish from various coastal locations in Europe or worldwide. Results are expressed in ng g1 ww unless otherwise stated. For PFCAs, compounds corresponding to maximum concentrations are specified in brackets. Location PFOS France (English Channel) France (Atlantic coast) France (Mediterranean coast) France (English Channel) France (Atlantic coast) France (Mediterranean coast) France (all coasts) France (Gironde estuary) France (all coasts) Spain (Catalonia) Mediterranean (Greek markets) Denmark Spain (North) Spain (North) Italy (Tyrrhenian sea) Baltic Sea, North Sea Spain (Catalonian markets) Portugal (North estuaries) UK Mediterranean Sea PFCAs France (English Channel) France (Atlantic coast) France (Mediterranean coast) France (English Channel) France (Atlantic coast) France (Mediterranean coast) France (Gironde estuary) Mediterranean (Greek markets) Spain (North) Korea Korea Korea Brazil Baltic Sea, North Sea China Korea China South China, Japan Ariake Sea, Japan

Sampling period

Concentrations min-max/median or mean

Organism

Reference

2016e2017 2016e2017 2016e2017 2013e2015 2013e2015 2013e2015 2011e2012 2012 2010 2011 2011 e 2009 2006e2009 2008 2006e2008 2006 e 2006 e

0.008e0.218/0.100 0.015e0.181/0.084 0.007e0.162/0.035 0.115e0.549/0.210 0.013e0.364/0.110 0.017e0.173/0.051 <0.056e0.681/0.271 0.085e0.15 0.005e0.874/0.081 2.70 <0.49
Mussel/Oyster Mussel/Oyster Mussel Mussel/Oyster Mussel/Oyster Mussel Mussel/oyster Oyster Mussel/oyster Fish and seafood Mussel Mussel Mussel (caged) Oyster (caged) Mussel Mussel Mussel, shrimp Mussel Oyster Mussel/Clam

This study This study This study This study This study This study Munschy et al. (2015) Munoz et al. (2017) Munschy et al. (2013) Domingo et al. (2012) Vassiliadou et al. (2015) Bossi et al. (2008)  mez et al., 2011 Go Fern andez-Sanjuan et al. (2010) Renzi et al. (2013) Rüdel et al., 2011 Ericson et al. (2008) Cunha et al. (2005) Clarke et al. (2010) Nania et al. (2009)

∑ PFCAsb ∑ PFCAsb ∑ PFCAsb ∑ PFCAsb ∑ PFCAsb ∑ PFCAsb C9 to C14 C4 to C16 PFOA, PFNA C4-C11 C5-C12

2016e2017 2016e2017 2016e2017 2013e2015 2013e2015 2013e2015 2012 2011 2009 2010 2009

Mussel/Oyster Mussel/Oyster Mussel Mussel/Oyster Mussel/Oyster Mussel Oyster Mussel Mussel (caged) Mussel, Oyster Bivalves

This study This study This study This study This study This study Munoz et al. (2017) Vassiliadou et al. (2015)  mez et al., 2011 Go Hong et al. (2015) Naile et al. (2012)

C4-C11 C7-C12 C5-C12 C7, C10, C11 C8-C12 C6-C11 C6-C12 C7, C8

2008 2008 2006e2008 2006e2007 2006 2004 2004 2003

0.036e0.491/0.147 0.067e0.493/0.180 0.150e2.374/0.401 0.017e0.479/0.222 0.014e0.429/0.165 0.078e1.979/0.357
Mussel, Mussel Mussel Mussel, Mussel, Mussel, Mussel, Mussel,

Naile et al. (2010) Quinete et al. (2009) Rüdel et al., 2011 Pan et al. (2010) Yoo et al. (2009) Gulkowska et al. (2006) So et al. (2006) Nakata et al. (2006)

Lines highlighted in bold refer to the present study. a Mean of means calculated from data presented in Cunha et al. (2005). b Sum of C6- to C14- PFCAs. c Recalculated from concentration in dw with a percentage of humidity of 80%.

Oyster

Oyster Oyster Oyster Oyster Oyster

M. Catherine et al. / Chemosphere 228 (2019) 640e648

Current data shows evidence of a global decrease in PFOS levels on French coasts over the years (Table 1). A PFOS decrease has also been observed in other European locations (see discussion, section 3.3), in line with regulations on PFOS direct production sources. Sampling years must therefore be taken into consideration and could represent confounding factors in data comparison. PFOS concentrations in European shellfish have been found to vary widely across locations, according to the characteristics of the sampling sites, potential local sources and study years. The highest concentrations were reported in mussels from industrialized estuaries in Portugal (extremely high values) and oysters from the UK (Cunha et al., 2005; Clarke et al., 2010). Other studies have reported concentrations in the same range as those found in our study. In particular, the levels reported by Munoz et al. (2017) in oysters from the Gironde estuary in France were very similar to ours. P PFCA concentrations on French coasts were globally of the same order of magnitude as those found in other shellfish worldwide (Table 1). However, huge differences in levels and profiles were found across study sites, as illustrated by the following examples. PFTrDA and PFTeDA - the two compounds detected most frequently and at the highest levels in our samples - are very seldom reported in shellfish in the literature (
3.3. PFAS profile temporal shifts P PFCA concentrations were higher than PFOS concentrations P (hence above PFSAs) in all samples collected in 2016e2017, i.e., a major shift versus our previous reports, in which PFOS was found to be the predominant PFAS in most samples collected from the English Channel and Atlantic coast, while PFCAs prevailed in Mediterranean samples (Munschy et al., 2015). PFOS represented 64% of total PFASs in samples from English Channel and Atlantic coast in P 2013e2014, dropping to 31% in 2016e2017. PFCAs were hence predominant at all sites. In addition, and similarly to previous observations (Munschy et al., 2015), Mediterranean samples were characterized by a significantly (p < 0.001) higher PFCAs/PFOS ratio (mean value of 20) than samples from the English Channel and Atlantic coast (mean value of 2) in 2016 and 2017. Interestingly, two sites showed dissimilarities in PFOS and PFCA relative profiles with regards to other samples from the same coastline. One site, located P on the Atlantic coast (the Nivelle estuary), showed a PFCAs/PFOS ratio of 24 and 26 in 2016 and 2017, respectively, which was similar to the Mediterranean samples and may reflect the influence of P specific local sources. Also, the PFCAs/PFOS ratio at one Mediterranean site (Thau lagoon: 3.6 on average in both years) was similar to that determined at sites in the English Channel and on P the Atlantic coast. These results show that the long-chain PFCA/ PFOS concentration ratio is site-specific and could therefore be efficient in pinpointing source spatial discrepancies across sites.

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3.4. Temporal trends in PFAS levels Time-series studies on emerging contaminants play an important role in assessing variations in emissions and evaluating the impact of legal restrictions on environmental levels. This study revealed that PFOS concentrations determined in samples collected in 2016 and 2017 had globally decreased at most sites versus our previous study (Munschy et al., 2015). PFOS median concentrations (across all sites) were 0.122 ng g1 ww in 2013e2015 versus 0.066 ng g1 ww in 2016e2017 (significantly different at p < 0.05). P Conversely, PFCA concentrations exhibited stable median concentrations (0.205 ng g1 ww at all sites) over the entire study period. Recent temporal variations in concentrations were compared to our previous study conducted using retrospective analyses of archived samples (Munschy et al., 2013, 2015). The updated results for the Seine estuary site are shown in Fig. 3. The significant downward trend in PFOS concentrations previously observed from the early 1990s until 2011 was confirmed in recent years (significant decrease at >99.9% confidence level, Table S1), although at a slower rate, with increasingly-stable concentrations from 2009 onwards. The levelling-off of PFOS concentrations observed in more recent years reflects environmental stabilization in response to P decreased inputs due to production bans. Conversely, PFCA concentrations were below LOQs prior to 2010 and showed a slight but significant (98.4% confidence) increase, with higher levels than PFOS from 2014 onwards. PFOS concentrations showed a similar time trend at the Gulf of Fos site in the Mediterranean Sea, with a significant linear decrease in levels after the mid-1990s (confidence value 99.9%). Conversely, P PFCAs showed increased levels after 2000, prior to which they were below LOQs (open squares in Fig. 4). No significant trends were found in the higher concentrations measured subsequently P and no global increase was indicated in PFCAs in the last 17 years P (61.9%). However, the overall PFCAs trend may mask the trends of individual compounds. For example, PFTeDA concentrations were on a probable uptrend in the Gulf of Fos (90.7%), while PFUnDA concentrations were on a significant downtrend (98%) (Fig. S1 and Table S1). Several studies have reported temporal trends in PFAS concentrations at various locations and matrices worldwide and shown conflicting results across sites and compounds. Further to bans, decreases are generally expected to be observed more rapidly at

P Fig. 3. Time trends of PFOS (black squares) and PFCA (open circles) concentrations 1 (ng g ww) in shellfish samples collected from 1981 to 2017 at the Seine estuary site (English Channel). Updated from Munschy et al. (2013). Dashed lines indicate signifiP cant PFOS trends. For PFCAs, the full line is a 2-point running mean smoother.

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P Fig. 4. PFOS (black squares, dash line indicates significant PFOS trends) and PFCA (black circles, open circles indicate LOQs; concentrations were < LOQs for the corresponding years) concentrations (ng g1 ww) in shellfish samples collected from 1981 to 2017 in the Gulf of Fos (Mediterranean coast).

t et al., 2013). Land et al. (2018) sites close to emission sources (Rige very recently published a thorough review and re-examination of previously-reported trends. Overall, the authors conclude that PFOS concentrations are not showing a global, post phase-out decline. In fact, in particular in European marine biota, various studies detected no significant trends in PFOS levels in marine mammals, fish or shellfish (Ahrens et al., 2009; Galatius et al., 2011; Kratzer et al., 2011; Rüdel et al., 2011; Huber et al., 2012; Ullah et al., 2014), while others detected a significant downtrend, as observed in eelpout from the North Sea between 2000 and 2009 (Rüdel et al., 2011) and shellfish from the Seine estuary, France, between the mid-1990s and 2011 (Munschy et al., 2013). Our current study shows that despite a levelling-off, a decrease in PFOS levels was confirmed with high confidence (>99.9%) until 2017, using a statistically-robust method. Regarding long-chain PFCAs, trends reported in the literature were either increasing (Galatius et al., 2011; Rüdel et al., 2011; Huber et al., 2012; Axmon et al., 2014) or stable (Rüdel et al., 2011; Huber et al., 2012). Significant trends were found to be opposite to those observed for PFOSs (Ahrens et al., t et al., 2013) and in agreement 2009; Houde et al., 2011; Rige with our findings.

4. Conclusions The results of our study show widespread occurrence of both PFOS and long-chain PFCAs in filter-feeding shellfish on French coasts. PFTrDA and PFTeDA were strongly correlated and pinpointed as the two predominant PFCAs at all sites. The highest PFOS and PFCA concentrations were determined in shellfish collected from main estuaries (Seine and Loire) and next to an industrial site (Gulf of Fos, Mediterranean coast), respectively. A decrease in PFOS concentrations was demonstrated across all study sites between P 2013-2015 and 2016e2017, while PFCAs globally showed similar P concentrations over the 5 study years. The PFCA/PFOS concentration ratio showed an increase over time, with site-specific values suggesting it could be used to track sources. The decreasing trend in PFOS concentrations previously detected (1990e2011) at the Seine estuary site in the English Channel was shown to be levelling-off, P while PFCAs showed a significant uptrend from 2010, in P contrast with the stable and higher levels of PFCAs found after 2000 in the industry-impacted site of the Gulf of Fos in the Mediterranean Sea. These trends are consistent with the progressive decrease in PFOS use and concurrent shift in production towards long-chain PFCAs. The data obtained will serve as a reference for future studies, either on a French or European scale.

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