### 2.4.3.2. Sample preparation procedure.
The purified extract was analyzed using liquid chromatography (Agilent Technologies, Santa Clara, CA, USA) coupled with tandem mass spectrometry. The separation of compounds was achieved using a Gemini C18 reverse phase column (3 µm, 50 × 2.0 mm) (Phenomenex, Torrance CA, USA). Elution solvents were MeOH 100% (A) and ammonium acetate in water (0.02 M) (B). The flow rate was 0.6 mL/min. The elution gradient was: t = 0–2 min: 30% A, t = 7–11 min: 100% A, t = 15–20 min: 30% A. The MS instrument used for detection and quantification was a triple quadrupole Agilent 6410 (Agilent Technologies, Santa Clara, CA, USA). The electrospray ionization was used in the negative mode. A minimum of two transitions were recorded per analyte. Mass spectra were acquired in the SRM mode. The common optimized source parameters were: capillary voltage, −3 kV; desolvation gas flow rate, 10 L/min; desolvation temperature was set at 300 °C; nebulization gas pressure at 45 psi and source temperature at 250 °C. The fragmentor voltage and collision energy were optimized for each monitored PFAS. Signals were treated by Agilent MassHunter Workstation Software and concentrations of PFAS are reported as ng/g wet weight (ww), ng/g dry weight (dw) and ng/g lipid weight (LW). The LOQ value was determined for each target compound in each analyzed sample. LOQ values were between 0.006 ng/g ww (PFOA) and 1.259 ng/g ww (PFBA) for each PFAS. The recoveries ranged from 65 to 125%.

### 2.4.3.3. LC-MS/MS measurement.
The purified extract was analyzed using liquid chromatography (Agilent Technologies, Santa Clara, CA, USA) coupled with tandem mass spectrometry. The separation of compounds was achieved using a Gemini C18 reverse phase column (3 µm, 50 × 2.0 mm) (Phenomenex, Torrance CA, USA). Elution solvents were MeOH 100% (A) and ammonium acetate in water (0.02 M) (B). The flow rate was 0.6 mL/min. The elution gradient was: t = 0–2 min: 30% A, t = 7–11 min: 100% A, t = 15–20 min: 30% A. The MS instrument used for detection and quantification was a triple quadrupole Agilent 6410 (Agilent Technologies, Santa Clara, CA, USA). The electrospray ionization was used in the negative mode. A minimum of two transitions were recorded per analyte. Mass spectra were acquired in the SRM mode. The common optimized source parameters were: capillary voltage, −3 kV; desolvation gas flow rate, 10 L/min; desolvation temperature was set at 300 °C; nebulization gas pressure at 45 psi and source temperature at 250 °C. The fragmentor voltage and collision energy were optimized for each monitored PFAS. Signals were treated by Agilent MassHunter Workstation Software and concentrations of PFAS are reported as ng/g wet weight (ww), ng/g dry weight (dw) and ng/g lipid weight (LW). The LOQ value was determined for each target compound in each analyzed sample. LOQ values were between 0.006 ng/g ww (PFOA) and 1.259 ng/g ww (PFBA) for each PFAS. The recoveries ranged from 65 to 125%.

### 2.4.4. Quality assurance/quality control

PCB, PBDE and PFAS concentrations were determined at LABERCA, the French National Reference Laboratory in charge inter alia of POPs in food and feed. All procedures were integrated in a certified quality assurance system ISO/IEC 17025:2005. A blank control and fish oil as internal quality control were included in each series of analysis. Results of this quality control sample were plotted in control charts to check a deviation. The compounds were quantified by isotopic dilution using the corresponding ¹³C-labeled isomers and internal standard method was used to quantify samples. The PCB procedure integrated quality control parameters to fulfill the requirements of Commission Directives 2002/69/EC and 2005/7/EC laying down the sampling methods and the methods of analysis for the official control of dl-PCBs in foodstuffs and feeding stuffs respectively. The laboratory applied the same requirements for PBDE analysis. The analytical method for PFAS was validated according to Commission Decision 2002/657/EC and 2004/25/EC concerning the performance of analytical methods and the interpretation of results. Furthermore, for each family of contaminants,

LABERCA successfully participates in international proficiency tests in food (Norwegian Institute of Public Health and Amsterdam University).

## 2.5. OH-PAH, AP and BPA analysis in fish bile

### 2.5.1. Chemicals and reagents

BPA, 5 OH-PAHs (2-naphtol (2-OHNa); 2-hydroxyfluorene (2-OHF); 1-phenanthrenol (1-OHPhe); 9-phenanthrenol (9-OHPhe); 1-hydroxypyrene(1-OHPyr)), 4 APs (4-tert-butylphenol (4-t-BP); 4-tert-octylphenol (4-t-OP); 4-octylphenol (4-OP); 4-n-nonylphenol (4-n-NP) (CAS #104-40-5), and a mixture of 4-p-nonylphenol (4-p-NP) (CAS #84852-15-3)) were purchased from Sigma Aldrich (France) as well as sodium acetate, ascorbic acid and β-glucuronidase with 1% sulfatase activity (Type H-3 from *Helix pomatia*). Ethanol (absolute) and acetonitrile (Scharlau LC-MS grade) were obtained from Intechim (France). To perform the analysis of total biliary proteins, sodium chloride, Trizma base and Bio-Rad Protein Assay were respectively purchased from Merck, Sigma-Aldrich (France) and Bio-Rad (France).

### 2.5.2. Sample preparation procedure

Conjugated metabolites of bile were analyzed according to modified methods from Kammann (2007) and Jonsson et al. (2008). Twenty-five µL of fish bile were incubated at 40 °C overnight in the presence of 100 µL of sodium acetate buffer (pH 5) and 25 µL of β-glucuronidase sulfatase (100 units β-glucuronidase and 5 units sulfatase/µL bile). After incubation, 100 µL of an ascorbic acid solution in ethanol (5 mg/L) were added and the mixture was vortexed for 1 min. A centrifugation was performed at 4 °C, 10,000 G during 10 min and the supernatant was removed and injected in the High Performance Liquid Chromatography (HPLC) system.

### 2.5.3. HPLC-FL measurements

The HPLC system was an LC Ultimate 3000 (Dionex – Thermo Scientific) equipped with a quaternary pump (DGP-3600RS), a thermally controlled auto-sampler (fixed at 4 °C) (WPS-300TSL), a column oven (TCC-3000SD) and a programmable fluorescence detector (FLD-3400RS, Dionex –Thermo Scientific). Chromatographic separation was performed on a reverse phase C18 column (5 µm, 250 mm × 3.0 mm; Supelcosil LC-PAH, Supelco, Sigma Aldrich) protected by a guard column (Supelgard LC-18, Supelco, Sigma-Aldrich) and maintained at 50 °C. Aliquots of 10 µL were injected and the flow rate was set at 1 mL/min. The mobile phase consisted of water (A): acetonitrile (B) with the following linear gradient: t = 0–5 min: 60% A; t = 15 min: 20% A; t = 20–23 min: 100% A; and t = 25–35 min: 60% A. Before analysis, the lambda max of excitation and emission of each targeted compound was determined using a 3D spectrofluorimeter (Perkin Elmer) and optimized with the fluorescence detector of the HPLC system used for the measurements (3D synchronous data acquisition). The excitation/emission wavelengths during the chromatographic run were: 225/313 nm for BPA and APs; 245/380 nm for 9-OHPhe, 1-OHPhe and 1-OHPyr; 271/334 nm for 2-OHF; 226/354 nm for 2-OHNa. The fluorescence detector allowed an analysis using simultaneously 4 channels. Data were acquired and processed using the Chromeleon 7 software (Thermo Scientific). The contaminant quantification was performed by an external calibration using multilevel calibration curves (at least 6 concentrations) from a mixture of APs (linear range: LOQ–150 ng/mL), BPA (linear range: LOQ–150 ng/mL) and PAH metabolites (linear range: LOQ–200 ng/mL).

### 2.5.4. Recovery and limit of detection

Concentrations of AP, BPA and PAH metabolites were determined at MMS laboratory of Nantes University. The sample preparation procedure was validated using spiked samples. The recoveries of APs, BPA and PAH metabolites (OH-PAHs) were calculated before and after extraction to evaluate extraction efficiency and matrix effects. A spike solution containing all compounds at a concentration of 100 ng/mL in
acetonitrile was added to fish bile randomly chosen. The spiked bile was treated as a sample. A blank sample (bile without spike solution) was also analyzed to determine background levels. These procedures were repeated three times. Recoveries obtained were: 92.4% for the BPA, 89.6% for the 4-t-BP, 89.0% for the 4-t-OP, 80.9% for the 4-OP, 79.9% for the 4-NP, 84.3% for the 2-OHNa, 79.2% for the 2-OHF, 85.9% for the 1-OHPhe, 47.6% for the 9-OHPhe and 89.0% for the 1-OHPyr.

The limits of detection and quantification (LOQs) were determined for each compound following the method described by Vial and Jardy (1999). The LOD values were: 0.25 ng/mL for the BPA, 0.18 ng/mL for the 4-t-BP, 0.41 ng/mL for the 4-t-OP, 0.003 ng/mL for the 4-OP, 0.46 ng/mL for the 4-NP, 0.017 ng/mL for the 2-OHNa, 0.033 ng/mL for the 2-OHF, 0.039 ng/mL for the 1-OHPhe, 0.48 ng/mL for the 9-OHPhe and 4.03 ng/mL for the 1-OHPyr.

Contents of APs, BPA and PAH metabolites were normalized by the measure of total biliary proteins and expressed as µg per g of protein. This normalization permitted the compensation of the differences of water content in the bile and the differences due to the feeding status of fish (Ruddock et al., 2003). The biliary protein content was measured by visible-spectrometry (UV-visible spectrophotometer Helios Omega – Thermo Scientific) using bovine γ-globulin as standard protein (Bio-Rad Protein Assay) following the method of Bradford (Bradford, 1976).

## 2.6. Statistical analysis

Data were analyzed using Statistica software (v 7, StatSoft, Inc.). The conditions for applying parametric tests, i.e. homogeneity of variance and normality, were verified using Bartlett and Shapiro-Wilk tests respectively. As result of these tests, non-parametric tests (Kruskal-Wallis ANOVA (KW)) were used in order to highlight significant differences of contaminant levels in the muscle and bile of eels from different locations and life stages. The KW test was followed by a post hoc test Multiple Comparisons of p-value (MCP) when it was significant at *p* < 0.05. Differences between groups were considered significant when *p* < 0.05. Links between variables were analyzed using the Pearson's correlation. As an integrated approach, Principal Component Analysis (PCA) was performed to illustrate and facilitate the discrimination of contaminant levels in the muscle and bile of eels from different sites and life stages, as well as their biometric parameters.

# 3. Results and discussion

## 3.1. Biometric parameters

The biometric parameters (Table 1) were presented according to the life stage (yellow or silver) of organisms and the fishing locations along the Loire estuary (Varades, Bellevue and Haute Indre). Four groups were considered: yellow eels from Varades (YV), from Bellevue (YB), from Haute Indre (YHI) and silver eels (SV). Total length mean was significantly greater for the YB group compared with YHI and YV (*p* = 0.002 and *p* < 0.001 respectively) but not significantly different to SV (*p* = 0.524). The lowest mean size was observed for YV which was very close to the YHI group. YHI and SV constituted intermediate groups. Nonetheless, SV had a greater mean size than YHI. Body weight differences between groups were similar to those found for the total length and showed the same significant differences. Whatever the life stage and the fishing location, BW was positively correlated to TL (*p* < 0.001). The linear regression equations were TL = 0.056BW + 39.62 R² = 0.887 for YV, TL = 0.071BW + 36.64 R² = 0.901 for YHI, TL = 0.024BW + 54.10 R² = 0.764 for YB and TL = 0.039BW + 46.71 R² = 0.946 for SV. These results are in accordance with a previous study on eels (*A. anguilla*) from the Loire estuary (Blanchet-Letrouvé et al., 2014). The three groups of yellow eels had a lower LW than silver eels but this difference was significant only for YV and YHI eels (*p* < 0.001 and *p* = 0.040,

respectively). These LW values were usually observed in silver European eels (Quadroni et al., 2013). Previous studies showed that yellow eels have lower lipid content mainly due to their growth whereas silver eels are in life development and accumulating lipid reserves for migration (Durif et al., 2005). According to Boëtius and Boëtius (1980), the percentage of lipids necessary to support the energetic cost of gamete production and migration must be at least 20% of the body weight. Only one SV eel had a lipid weight lower than this value (LW = 18.8%). All other silver eels had a lipid content ranging from 22.2% to 34.8%. Two more recent studies (Van den Thillart et al. (2004); Palstra et al. (2007)) established at 15% and 13.5% of the body weight, respectively, the lipid content necessary to realize the healthy migration to the Sargasso Sea. In the present study, it can be assumed the silver eels had sufficient lipid energy required for a successful reproductive migration.

The range values of GSI for the SV group were 1.66 to 2.21% reflecting that all individuals of this group were in the silver stage. The GSI values of yellow individuals were significantly lower (*p* < 0.008) and ranged from 0.11 to 1.62%. Only three yellow individuals out of 30 showed GSI values higher than 1.2% which could indicate the onset of silvering. YB and YHI tend to present higher values of GSI compared with YV but the difference was not significant. No sampling site effect was evidenced on the liver weight of yellow eels, HSI values being very variable and ranging from 0.79 to 1.61. YB eels showed significantly lower HSI values compared with SV (*p* < 0.05) but this difference was not found with other yellow eel groups. Durif et al. (2005) described a higher mean HSI value for yellow eels (1.58 ± 0.53) coming from different French waters and sampled between 1994 and 2002.

Silver eels were less infested by *Anguillcola crassa* than the yellow eels as showed by the PI. YHI presented the highest mean of PI among yellow individuals but the variability intra-group was high. The percentages of parasite eels (40%:YV; 66.7%: SV; 80%:YB and YHI) were in accordance with values usually observed in most eels from European habitats (40% up to 90%) (Palstra et al., 2007).

The Fulton's index constituted a biologic factor which could especially characterize the habitat context including pollution pressure and trophic conditions. On the whole, the Fulton's indexes of sampled eels were in the same range than those reported for eels fished in the Loire estuary in 2011 (Blanchet-Letrouvé et al., 2014) and in other European habitats (Bultel et al., 2014; Palstra et al., 2007; Reckordt et al., 2014; Tapie et al., 2011).

## 3.2. Contaminants in eel muscles

### 3.2.1. PCBs

The mean concentrations of dl-PCBs and PCBi according to the life stage and location in the estuary are presented in the Table 2. The 18 PCBs investigated were detected in all eels sampled in this study. The mean sum of these 6 PCBi for the yellow eels was 3050.7 ng/g LW. This concentration was lower than those measured in Flemish eels (Belpaire et al., 2011; Van Ael et al., 2014). In order to compare these results with previous studies, the PCB 118 was added to this sum. For yellow eels, the mean sum of the 6 PCBi + PCB-118 was twice lower (3326.9 ng/g LW) than the concentrations measured in the Rhin and Meuse (5724-8518 ng/g LW) by de Boer et al. (2010) and in Flanders (4636 ng/g LW) by Van Ael et al. (2014). However, this value was greater than levels reported in Italia (Tvere river: 685.3 ng/g LW Bettinetti et al. (2011)), in France (the Gironde estuary: 2062 ng/g LW, (Bodin et al., 2014), and in the Adour basin: 193-1320 ng/g LW (Tabouret et al., 2011)). However, Tapie et al. (2011) found, in eels with the same size class, close concentrations in the Gironde estuary (1913-4473 ng/g LW), the Garonne river (1372-3478 ng/g LW) and the Dordogne river (1321-2992 ng/g LW). PCB eel contaminations were previously investigated in the Loire estuary by Blanchet-Letrouvé et al. (2014). Two sites were directly comparable with this study because of their proximity (Varades and Nantes/Bellevue). The results
<table><caption>**Table 2**
Mean concentrations of dl-PCBs, ndl-PCBs (PCBi), PCBs (dl-PCBs + ndl-PCBs), PBDEs and PFAS in eel muscles and concentrations of BPA, APs and OH-PAHs in eel bile from Loire estuary according to the sampling site (Varades, Bellevue, Haute Indre) and the life stage (yellow and silver). Muscle concentrations are expressed in ng/g wet weight (ww), ng/g dry weight (dw) and ng/g lipid weight (LW) and bile concentrations are expressed in ng/mL bile and µg/g protein (prot.).</caption><thead><tr><th>Life stage</th><th></th><th>Yellow eels</th><th></th><th>Yellow eels</th><th></th><th>Yellow eels</th><th></th><th>Silver eels</th></tr><tr><th>Sampling site</th><th></th><th>Varades</th><th></th><th>Bellevue</th><th></th><th>Haute Indre</th><th></th><th>Estuary</th></tr><tr><th>n</th><th></th><th>10</th><th></th><th>10</th><th></th><th>10</th><th></th><th>15</th></tr></thead><tbody><tr><td rowspan="15">Muscle concentration</td><td>dl-PCB sum (ng/g ww)</td><td>10.9 ± 5.1</td><td>(4.6-21.3)</td><td>51.1 ± 25.0</td><td>(7.5-85.1)</td><td>38.8 ± 26.5</td><td>(7.9-89.7)</td><td>34.8 ± 7.0</td><td>(26.2-48.5)</td></tr><tr><td>dl-PCB sum (ng/g dw)</td><td>38.6 ± 14.2</td><td>(21.0-61.1)</td><td>143.8 ± 71.3</td><td>(16.1-277.4)</td><td>105.4 ± 50.3</td><td>(34.5-201.9)</td><td>78.9 ± 16.7</td><td>(60.2-106.4)</td></tr><tr><td>dl-PCB sum (ng/g LW)</td><td>343.7 ± 350.4</td><td>(88.0-1156.1)</td><td>566.7 ± 701.4</td><td>(25.3-2418.4)</td><td>506.9 ± 374.0</td><td>(176.4-1225.3)</td><td>132.8 ± 33.9</td><td>(89.1-208.0)</td></tr><tr><td>ndl-PCB sum (ng/ww)</td><td>63.4 ± 28.2</td><td>(29.1-125.2)</td><td>329.4 ± 207.7</td><td>(22.5-746.2)</td><td>256.2 ± 170.2</td><td>(69.3-635.8)</td><td>190.0 ± 34.7</td><td>(144.1-253.5)</td></tr><tr><td>ndl-PCB sum (ng/g dw)</td><td>223.6 ± 70.0</td><td>(131.7-359.4)</td><td>916.9 ± 538.3</td><td>(48.5-2005.4)</td><td>717.0 ± 342.0</td><td>(304.1-1430.9)</td><td>430.8 ± 86.9</td><td>(345.3-607.3)</td></tr><tr><td>ndl PCB sum (ng/g LW)</td><td>1945.3 ± 1922.2</td><td>(587.5-6659.9)</td><td>3355.7 ± 3583.9</td><td>(76.1-12270.2)</td><td>3851.2 ± 3367.2</td><td>(782.4-10795.5)</td><td>724.6 ± 180.7</td><td>(533.2-1078.5)</td></tr><tr><td>PCB sum (ng/g ww)</td><td>74.3 ± 33.2</td><td>(33.8-146.5)</td><td>380.5 ± 230.3</td><td>(30.0-821.5)</td><td>294.9 ± 195.7</td><td>(77.2-725.5)</td><td>224.8 ± 41.2</td><td>(172.2-302.0)</td></tr><tr><td>PCB sum (ng/g dw)</td><td>262.2 ± 83.9</td><td>(152.7-420.6)</td><td>1060.7 ± 600.7</td><td>(64.6-2207.7)</td><td>822.4 ± 389.8</td><td>(338.6-1632.8)</td><td>509.8 ± 102.5</td><td>(413.6-702.7)</td></tr><tr><td>PCB sum (ng/g LW)</td><td>2288.9 ± 2270.8</td><td>(675.5-7816.1)</td><td>3922.5 ± 4276.4</td><td>(101.5-14688.6)</td><td>4358.1 ± 3740.2</td><td>(958.8-12020.8)</td><td>857.4 ± 213.1</td><td>(622.3-1286.5)</td></tr><tr><td>PBDE sum (ng/g ww)</td><td>0.6 ± 0.5</td><td>(0.1-1.9)</td><td>4.9 ± 3.1</td><td>(1.8-10.9)</td><td>4.2 ± 3.2</td><td>(0.7-8.8)</td><td>9.5 ± 2.9</td><td>(6.7-18.1)</td></tr><tr><td>PBDE sum (ng/g dw)</td><td>2.2 ± 2.3</td><td>(0.4-8.1)</td><td>12.9 ± 6.1</td><td>(6.4-25.0)</td><td>10.9 ± 6.0</td><td>(2.9-18.0)</td><td>21.6 ± 6.4</td><td>(15.4-39.9)</td></tr><tr><td>PBDE sum (ng/g LW)</td><td>25.9 ± 42.1</td><td>(2.3-107.2)</td><td>40.5 ± 29.7</td><td>(10.0-102.9)</td><td>46.2 ± 29.2</td><td>(24.1-103.6)</td><td>36.3 ± 12.0</td><td>(21.8-70.0)</td></tr><tr><td>PBDE 47 (ng/g ww)</td><td>0.3 ± 0.3</td><td>(0.0-1.1)</td><td>3.0 ± 2.0</td><td>(0.9-6.7)</td><td>2.4 ± 2.1</td><td>(0.2-5.6)</td><td>6.5 ± 2.3</td><td>(4.2-13.5)</td></tr><tr><td>PBDE 47 (ng/g dw)</td><td>1.2 ± 1.3</td><td>(0.2-4.6)</td><td>7.8 ± 4.0</td><td>(3.7-15.3)</td><td>6.1 ± 4.0</td><td>(0.9-11.4)</td><td>14.6 ± 5.0</td><td>(10.1-29.7)</td></tr><tr><td>PBDE 47 (ng/g LW)</td><td>13.1 ± 20.7</td><td>(1.3-60.6)</td><td>22.4 ± 12.7</td><td>(7.5-47.4)</td><td>21.6 ± 8.3</td><td>(10.1-36.0)</td><td>24.6 ± 9.3</td><td>(14.4-52.0)</td></tr><tr><td>PFAS sum (ng/g ww)</td><td>22.5 ± 16.0</td><td>(5.1-47.8)</td><td>45.2 ± 44.1</td><td>(7.7-149.3)</td><td>20.0 ± 16.9</td><td>(4.1-50.4)</td><td>32.5 ± 23.4</td><td>(5.1-84.0)</td></tr><tr><td>PFAS sum (ng/g dw)</td><td>88.4 ± 71.9</td><td>(16.4-212.7)</td><td>123.7 ± 104.6</td><td>(16.5-342.1)</td><td>75.3 ± 76.2</td><td>(8.7-221.0)</td><td>75.4 ± 57.6</td><td>(12.0-210.0)</td></tr><tr><td>PFAS sum (ng/g LW)</td><td>1293.0 ± 2090.3</td><td>(50.5-6656.8)</td><td>398.5 ± 384.1</td><td>(26.0-1256.2)</td><td>1177.8 ± 2434.9</td><td>(12.4-7845.9)</td><td>129.9 ± 104.2</td><td>(18.7-370.5)</td></tr><tr><td rowspan="6">Bile concentration</td><td>BPA (ng/mL)</td><td>3109.1 ± 1706.5</td><td>(1715.5-7012.4)</td><td>412.7 ± 705.2</td><td>(15.5-2321.2)</td><td>&lt;LOD</td><td>-</td><td>36.8 ± 34.2</td><td>(18.6-125.9)</td></tr><tr><td>BPA (µg/g prot.)</td><td>1212.9 ± 1278.7</td><td>(147.0-3641.5)</td><td>127.5 ± 184.0</td><td>(8.4-564.9)</td><td>&lt;LOD</td><td>-</td><td>5.2 ± 5.2</td><td>(1.7-15.6)</td></tr><tr><td>AP sum (ng/mL)</td><td>189.0 ± 107.1</td><td>(46.8-349.3)</td><td>141.0 ± 114.7</td><td>(35.3-358.1)</td><td>617.8 ± 237.8</td><td>(141.8-842.7)</td><td>160.8 ± 232.9</td><td>(0.0-569.2)</td></tr><tr><td>AP sum (µg/g prot.)</td><td>59.3 ± 39.5</td><td>(16.6-125.1)</td><td>64.2 ± 51.9</td><td>(21.4-170.0)</td><td>285.5 ± 226.1</td><td>(33.3-575.6)</td><td>30.7 ± 30.6</td><td>(4.4-70.4)</td></tr><tr><td>OH-PAH sum (ng/mL)</td><td>74.5 ± 44.3</td><td>(35.0-167.5)</td><td>23.9 ± 20.4</td><td>(0.5-53.7)</td><td>32.6 ± 11.7</td><td>(15.9-50.8)</td><td>148.8 ± 68.1</td><td>(0.0-259.4)</td></tr><tr><td>OH-PAH sum (µg/g prot.)</td><td>26.0 ± 16.4</td><td>(3.7-43.7)</td><td>8.6 ± 10.0</td><td>(0.3-25.5)</td><td>14.9 ± 10.7</td><td>(4.6-32.8)</td><td>21.7 ± 6.3</td><td>(12.9-32.1)</td></tr></tbody></table>

Mean ± standard deviation; min-max value are presented into brackets.
were similar for the yellow eels from these sites as well as those of silver eels. The downstream site in this study (Haute Indre) was further away than the Cordemais site, leading to differences of eel contaminations. YHI eels seemed to be more contaminated than eels from the Cordemais site because the Haute Indre site is closer to Nantes city. The results of the present study showed intermediate contamination of yellow eels from the Loire estuary compared with other French rivers. Silver eels were less contaminated (sum of the 6 PCBi and 6 PCBi + PCB-118, 725 and 799 ng/g respectively) compared with the whole of the yellow ones, regardless of the sampling site. Regarding the influence of the fishing zone in the Loire estuary, the yellow eels from Varades (YV) were significantly (p < 0.027) less contaminated in ndl-PCBs and dl-PCBs expressed as ng/g ww and dw than the ones from the two other sites (YHI, YB). No differences were found between the YB and YHI. These results could be explained by the influence of the sampling site and the LW of eels. The Varades site is far from industrial and urban influences compared with the other sampling sites (Bellevue, Haute Indre). On the other hand, assuming that PCBs are lipophilic compounds, their levels are usually correlated to LW. In the present study, PCB levels (expressed in ww) were correlated to the LW only for YV and YHI groups (p < 0.05). The differences observed in PCB accumulation could be partially linked to lower LW depicted mainly for YV and to a lesser extent for YHI. Normalization of the PCB concentrations by eel lipid contents was more relevant for the comparison between the sampling sites, since other factors like the size and physiological condition also affected the overall bioaccumulation. This lipid normalization leads to similar levels whatever the sampling site, indicating that the contamination level in Varades was comparable with those in Bellevue and Haute Indre. However, the high variability of PCB concentrations in the different groups must be duly noted. This variability could be due to the fact that in estuarine habitats, furthermore in estuarine tidal systems, the sedentarity is less pronounced than in freshwater upstream habitats.

An influence of the life stage is only observable between YV and SV which was more contaminated (p < 0.038; for dl-PCB, PCBi and PCB sum in ww and dw). When concentrations of ndl-PCBs, PCBi and total PCBs were normalized by gram of lipids in muscle, YV levels were similar to SV and significantly (p < 0.005) lower than YB and YHI. These results could be explained by the fact that migrating silver eels could come from less polluted freshwater areas, far upstream,

since the Loire River is characterized by the absence of barriers permitting the free downstream-upstream migration of fish. Furthermore, Daverat et al. (2004) reported that eels would not be strictly sedentary fish. Moreover, lipophilic compounds like PCBs have the particularity of accumulating in fat muscle. Another hypothesis could be that the mobilization of lipid reserves would induce a detoxification process of the muscles of silver eels. In fact, silver eels during migration mobilized lipid reserves providing the necessary energy for sexual maturation and also for crossing the Atlantic Ocean to reach spawning sites (Durif et al., 2005). Since the Loire is a large river, a potential migration could start from areas located hundreds of kilometers upstream.

In all sampled eels, dominant congeners among the 18 analyzed PCBs were PCB 153 (39.7%), PCB 138 (23.7%), PCB 180 (14.6%) and PCB 118 (8.4%) such as commonly found in percentages in Europe, e.g. in Flanders (Van Ael et al., 2014) or Gironde estuary (Bodin et al., 2014). The 6 PCBi represented 85.9% of the total PCB concentration in yellow eel muscles and 84.5% in the silver eels. The 6 PCBi + PCB-118 contributed to 94.2% and 93.3% of the total PCB levels for the yellow and silver eels respectively. According to PCB relative composition in commercial mixtures, this proportional dominance of PCB 153, 138 and 180 could reflect the historic use and subsequent environmental dispersion of Aroclor 1260 (Belpaire et al., 2011; Frame et al., 2001). The patterns of the 6 PCBi + PCB-118 according to sampling site and life stage are illustrated in Fig. 2. No differences were found for the PCB congeners 28, 52, 101 and 138 (p > 0.05) between the eels of the 3 sampling sites. However, differences were found for the congeners 118, 153 and 180. The proportion of the PCB 153 increased for the eels fished in the downstream part of the river, notably for the YHI eels, which were significantly different than the YV eels (p < 0.001). Conversely, the proportion of the PCB 180 decreased in eels fished downstream in the estuary. Indeed, the level was lower in YHI eels than in YV eels (p = 0.004). For the PCB 118, the percentage was greater in the muscles of YB eels than in YHI eels (p = 0.041) whereas no significant difference was found between YV eels and the two other groups. In this area of the estuary, numerous variations in geochemical processes such as salinity, turbidity, temperature and tidal cycles could affect the bioavailability of contaminants such as PCBs between the different sampling sites (Blanchet-Letrouvé et al., 2014; Cailleaud et al., 2009; Turner, 2003). The results on PCB 153 and 180 were in accordance

<figure><img src="image_57.png" /><figcaption>**Fig. 2.** Profiles of the 6 PCBi + PCB-118 in yellow European eels from three monitored sites in the Loire estuary (YV: Varades (n = 10); YB: Bellevue (n = 10); YHI: Haute Indre (n = 10)) and silver eels (SV; n = 15). PCB proportions were calculated considering the sum of 6 PCBi + PCB-118. Error bars correspond to standard deviation. Error bars correspond to standard deviation.</figcaption></figure>
<figure><img src="image_58.png" /><figcaption>**Fig. 3.** Eel quality classes distribution in yellow European eels from three monitored sites in the Loire estuary (YV: Varades (n = 10); YB: Bellevue (n = 10); YHI: Haute Indre (n = 10)) and silver eels (SV; n = 15) according to 6 PCBi (ww).</figcaption></figure>

with those of Blanchet-Letrouvé et al. (2014) for Bellevue and Varades sites. The PCB 153 was considered as a non-metabolizable congener, i.e. a tracer of bioaccumulation process. Its higher proportion in YHI eels could suggest differences of diet, ecology, physiology or metabolization capacities, as already proposed by Tapie et al. (2011).

Regarding the influence of the life stage, SV eels presented similar patterns as the YV and YB eels, (except for the congener 28 which was significantly higher for the SV eels (p = 0.009)). The SV eels had greater proportions of the congeners 28, 118 and 180 than YHI and a lower proportion of the congener 153. These differences could be explained by the unknown origin of the SV eels. Moreover, the mobilization of the lipids during the silvering process could modify PCB patterns. During starvation, the neutral muscle lipids are more quickly metabolized than the polar ones for energy production (Osako et al., 2003).

According to all the data available for the Loire eels, the existing scoring systems (quality classes; Amilhat et al., 2013; Belpaire and Goemans, 2007; ICES, 2012) were applied to estimate the overall quality of the eels (Fig. 3). The eels were classified as not impacted, slightly impacted, impacted and strongly impacted with PCBi concentrations in eels <62, 62–155, 155–391, >391 ng/g ww respectively. This figure confirmed the higher contamination of yellow eels from Bellevue and Haute Indre, compared with Varades where more than a half of eels appeared to be not impacted. The silver eels showed less variability than the yellow ones and appeared in the whole as impacted.

### 3.2.2. PBDEs

PBDE accumulation in the eels is shown in Table 2. Among the 7 investigated PBDEs, 6 were found in all fish. The PBDE 183 was found in all SV but it was detected in 10 to 50% of yellow eels. Burreau et al. (2006) indicated a link between biomagnification and the bromination degree. The PBDEs with up to 6–7 bromine atoms, except congener 183, biomagnify while heavier PBDEs do not. It could partially explain why it was not found in all fish. The presence of this congener in all SV could indicate different exposure windows and sources compared with the yellow eels. The sum of PBDE concentrations ranged from 25.9 to 46.2 ng/g LW in yellow eels and was 36.3 ng/g LW in silver eels. Urbanized sites seemed to be characterized by greater PBDE concentrations than by rural sites as reported by Macgregor et al. (2010) or Roosens et al. (2010). All analyzed eels exceeded the EU EQS biota for the sum of six PBDE congeners (28, 47, 99, 100, 153 and 154): 0.0085 ng/g fish (European Commission, 2008, 2013). The levels of PBDEs varied considerably, ranging from 10 to around 2100 times above the EQS.

The comparisons of monitoring studies are difficult since the number of analyzed PBDE congeners varied. Indeed, there are no PBDE indicators, such as for PCBs, and the regulatory measures of the PBDE

mixture uses tend to stabilize or decrease the contamination levels. This is why, because of its abundance, PBDE 47 is sometimes used to represent PBDE contamination and therefore, it is easier to compare studies using levels of this congener.

In the silver eels from the river Rhine, mean PBDE concentration was 88.7 ng/g LW (PBDE 47 = 59.2 ng/g LW) and in the river Elbe, they were 33.5 ng/g LW and 30.2 ng/g LW for the yellow and silver eels respectively (sum of 9 PBDEs) (22.5 ng/g LW and 21.5 ng/g LW for PBDE 47) (Sühring et al., 2013). Roosens et al. (2010) reported PBDE levels ranging from 10 to 5811 ng/g LW for eels from Flanders (Σ of 10 PBDEs). This high variability was due to a large data set with data from many sites of different typology, area or river basin including healthy sites and sites within proximity of industrialized areas (textile, production of electronic devices and plastic industry). Recently, Malarvannan et al. (2014) presented PBDE values ranging from 12 to 1400 ng/g LW (Σ of 14 PBDEs) in eels from Flanders. Concerning eels from French rivers, Bragigand et al. (2006) found PBDE concentrations in eel muscles ranging from 3.5 to 13.1 ng/g ww in the Seine estuary and from 0.3 to 0.9 ng/g ww in the Loire (Σ6 of PBDEs) (54 ng/g LW and 1.6 ng/g LW for the PBDE 47). The eels from the Seine estuary were represented by yellow individuals and appeared to be more contaminated than the yellow eels in our study. The levels found in the eels from the Loire were obtained in the muscles of silver eels and were lower in comparison with our data. A study on eels of same size class from the Gironde estuary reported concentrations between 61 and 237 ng/g LW (Tapie et al., 2011). Compared with the PBDE concentrations found in eels from the European waters, the eels from the Loire estuary, like those from the Irish waters (McHugh et al., 2010) and from the French Mediterranean coastal lagoons (Labadie et al., 2010), seemed to be less contaminated. Since there are few companies using PBDEs in the basin of the Loire estuary (Bragigand, 2005), moderate PBDE levels in the Loire should mostly result from the use and incineration of goods commonly consumed in everyday life (Bragigand et al., 2006). Furthermore, penta-BDE and octa-BDE mixtures have been banned in Europe since 2004 and the deca-BDE, the only PBDE mixture authorized, is no longer produced in France or in the European Union (INERIS, 2012), contributing as well to a moderate contamination. Urbanized sites tended to have greater PBDE levels than rural sites as reported for example by Macgregor et al. (2010) or Roosens et al. (2010).

Among the three sites studied in the Loire estuary, YV eels showed PBDE concentrations (expressed in wet and dry weights) significantly lower than the YB and YHI eels (p < 0.05). The YHI eels downstream the river showed PBDE levels (expressed in LW) twice greater than YV eels (p = 0.036), YB eels being intermediate. PBDE contaminations would be increasing in the direction of the mouth of the estuary. YB and YHI eels located in urbanized and industrialized sites seem to be greater exposed to PBDEs than the YV eels under agricultural pressure.

According to the life stage, silver eels were significantly more contaminated by PBDE compounds than the yellow eels (p < 0.036), except for the YB eels. YB eels presented singular physiological characteristics (BS, BW and LW) close to the SV ones which could explain this similarity. After the normalization of concentrations by the lipid content, all differences were annihilated suggesting correlation between PBDE concentration and lipid content. However, PBDE concentrations were significantly correlated with lipid content only for the YHI eels. The absence of correlation between PBDE concentration and LW for the SV eels, as well as for PCB levels, could be explained by the lipid mobilization and by the lack of information on their upstream habitat, which may be less contaminated than the estuary.

PBDE patterns are presented in Fig. 4. Whatever the life stage and location, the most abundant congeners in eel muscles were the PBDE 47 (59.9 ± 9.8%) followed by the PBDE 100 (30.7 ± 11.1%) and the PBDE 154 (5.2 ± 3.1%). PBDE 47 is usually the most abundant congener in fish due to its high uptake rate and to its high bioaccumulation factor (Andersson et al., 1999; Burreau et al., 2006). The high contribution of congener 47 in the patterns may also be due to the enzymatic
<figure><img src="image_59.png" /><figcaption>**Fig. 4.** Profiles of the 7 PBDEs in yellow European eels from three monitored sites in the Loire estuary (YV: Varades (n = 10); YB: Bellevue (n = 10); YHI: Haute Indre (n = 10)) and silver eels (SV; n = 15). PBDE proportions were calculated considering the sum of 7 PBDEs. Error bars correspond to standard deviation.</figcaption></figure>

debromination of higher brominated PBDEs in lower brominated ones, like 47, during the metabolization in fish (Eljarrat and Barceló, 2011). For the other PBDE analyzed congeners, the concentrations were low and sometimes below the analytical limit of detection, notably for the congener 183. The low proportion of PBDE 99 could be explained by three hypotheses: i) the nature of source contamination; ii) the low assimilation efficiency (Burreau et al., 1997; Stapleton et al., 2004a, 2004b); iii) the possible biotransformation/metabolism to PBDE 47 in eels as hypothesized by several authors (Labadie et al., 2010; Malarvannan et al., 2014; McHugh et al., 2010; Roosens et al., 2010). Such metabolic pathways have been described in carp (*Cyprinus carpio*) by Stapleton et al. (2004c). Analysis of spatial differences showed no significant differences between sites considering the PBDE 47 as well as the congeners 153, 154 and 183. The congeners 28 and 99 presented a significantly higher percentage in the YV eels than in the YB eels (p < 0.008). The YHI eels had a higher proportion of the PBDE 100 than the YV eels (p < 0.031).

Regarding life stage, differences were founded for 5 PBDE proportions but without evident coherence. Thus, the congener 28 concentration was lower in the SV eels than in the YV eels (p < 0.004). For the PBDE 47, the SV eels had a higher percentage than in YV and YHI eels (p < 0.013). The proportion of congener 100 was lower in the SV eels than in the YHI eels (p < 0.002). PBDE 153 had a higher proportion in the SV eels than in the YB eels (p < 0.007). And PBDE 183 were more represented in the SV eels than in the YB and YHI eels (p < 0.003). It must be duly noted that a greater proportion was observed in the pattern of the SV eels for the higher brominated congeners. The PBDE profile of SV eels consisted of PBDE 183 and PBDE 153, which are congeners of the technical octa-BDE mixture. Since August 2004, the penta-and octa-BDE mixtures are banned in Europe due to European scientific assessments (INERIS, 2012). Assuming that the silvering metamorphosis occurs after the juvenile growth and differentiation phase (>7 years for females) in the inland waters (Van Ginneken et al., 2007), their presence in the SV eels could probably be linked to the exposure to this technical mixture before the ban. It could also be induced by debromination to lower brominated congeners of the technical deca-BDE mixture as described by Eljarrat and Barceló (2011), this mixture being the only one authorized in Europe.

### 3.2.3. PFAS

The PFAS sums observed in eels from the estuary are presented in Table 2. The values were between 75.25 and 123.67 ng/g dw with great variations within the different groups. A total of 17 PFAS were investigated and only 2 (PFOS and PFUnA) were detected in all muscle eels. The PFOSi was not detected in any eels. Among the 5 perfluoroalkylsulfonate substances (PFASS), only the PFOS was found in all eels. In contrast, the PFBS was not detected in any of the samples. The PFOS is known for its high bioaccumulative potential, whereas PFBS is considered non-bioaccumulative (Conder et al., 2008), which is consistent with present results. The PFHxS was observed in all the yellow eels but not detected in 2 SV eels. PFHpS was present in about one quarter of individuals, mainly in the SV and YB eels, but never in all the fish of these groups. Only 8 out of the 11 studied perfluorocarboxylic acids (PFCAs) were detected in eel muscles. The PFHxA was in only one fish (YV) and the PFUnA was in all eels. Not detected PFCAs were the PFBA, the PFPA and the PFHpA. The 6 others were found in all the eel groups but not in all the fish. According to Conder et al. (2008), the PFCAs with shorter carbon chain possess a very low or non-existent bio-accumulation potential that could explain the lack of detection of the PFCAs with 7 or less carbons.

PFAS regulation is recent and involves only few compounds. Since May 2009, perfluorooctane sulfonic acid (PFOS), its salts and perfluorooctanesulfonyl fluoride (PFOSF) are part of the 9 substances added to the list of substances covered by the Stockholm Convention on Persistent Organic Pollutants (POPs) (UNEP, 2009). As such, PFOS and derivatives are covered (restrictions on the marketing and use for substances) by Regulation (EU) No 207/2011 of Commission Regulation and Directive (EU) 2006/122/ECOF of the European Parliament and of the Council of 2 March 2011 and 12 December 2006, respectively. The EU EQS specific to PFOS was 9.1 ng/g ww for biota (European Commission, 2013). Three quarters of eels fished in this study exceeded this value (YV: 70%, YB: 90%, YHI: 50% and SV: 73%). The highest value was approximately 14 times higher than the UE EQS PFOS and was found in the muscle of one YB eel. To our best knowledge, data on the eel contamination by PFAS in Europe are scarce. It is therefore difficult to compare the levels at the national and the European scales. The predominant PFAS found in all species, tissues and locations around the world is the
PFOS (Houde et al., 2011). The level of PFOS in the present study ranged from 67.6 to 106.9 ng/g dw (17.9 to 39.0 ng/g ww) for the yellow eels and was 61.6 ng/g dw (26.6 ng/g ww) for the silver ones. In German water, the average levels of PFOS in different fish (herring, mackerel, eel, trout, carp, bream) ranged from 12.0 to 135.7 ng/g ww and the compound was not detected in 3 sites (Lake Storkow, North and Baltic Seas) (Schuetze et al., 2010). In this same study, the PFOA was not detected in any investigated samples. The distribution of 15 perfluorinated compounds in eels from aquatic systems in The Netherlands was studied by Kwadijk et al. (2010). The authors found only 3 PFAS in all eel filets. The PFOS was found to be the predominant compound with concentrations ranging from 7 to 58 ng/g ww. The two other PFAS, the PFHxS and PFDoA, were detected at a level approximately 10 times lower than the PFOS. In a comparative earlier study conducted also in the Netherlands, the PFOS ranged between <8 to 143 ng/g ww (Schrap et al., 2004). Eels caught in Lake Möhne in Germany had PFOS concentrations ranging from 37 to 83 ng/g ww and a maximum PFOA concentration of 2.3 ng/g ww (Hölzer et al., 2011). However, the PFOA was not found in all fish from this lake and the other analyzed PFAS (PFDoA, PFUnA, PFDA, PFNA, PFHpA, PFHxA, PFHxS and PFBS) were not detected in any fish sample. Again, the absence or the lower levels of PFAS with a shorter carbon (<7 C) were found.

The presence of PFAS and particularly of PFOS in eels from the Loire could be explained by the industrial activities around the estuary. In fact, the Pays de la Loire region is France's foremost region for wood transformation, especially paper industry and for metal surface treatment industry, both of which represent potential sources of PFOS (Munschy et al., 2013).

Regarding the spatial (YV, YHI and YB) and the life stage (yellows vs silvers) differences, no significant differences between the PFAS sums were observed whatever the expression unit (ww, dw or LW) due to the very high variability of the concentrations of these groups, particularly expressed in LW. The PFAS accumulations in eels seem to be independent to the sampling site and life stage. In this study, the lipid content was negatively correlated with the PFAS sums, expressed as dw and ww, for the YHI eels (p < 0.05) and no correlation was found for the other groups.

The PFAS patterns (Fig. 5) showed that PFOS is the major detected PFAS. Its contribution to the sum was comprised between 84 and 88%. This is the major PFAS found in fish (Hoff et al., 2005; Hölzer et al.,

2011; Houde et al., 2011; Kwadijk et al., 2010; Labadie and Chevreuil, 2011; Miège et al., 2012; Schuetze et al., 2010). No differences between the three studied sites were found for the main PFAS (PFOS, PFDA, PFUnA and PFDoA). However, the YHI eels had a greater proportion of PFHxS than the two other groups. Regarding the life stage, the SV eels, like the YV and YB, had a lower proportion of PFHxS than YHI eels (p = 0.037). The YHI eels were caught after Nantes city thus the PFHxS contamination could be due to urban pressure (sewage effluents) and to industrial activities around this city, especially the paper industry. The PFUnA proportion was higher in the SV eels than in the YV eels (p = 0.048) among the main PFAS (PFOS, PFDA, PFUnA and PFDoA). The origin of this compound preferentially found in the SV eels was difficult to explain, notably due to the unknown living area of these silver individuals.

### 3.3. OH-PAHs, APs and BPA in eel biles

The BPA level, AP sums and OH-PAH sums are presented in Table 2. The 4-OP and the 9-OHPhe were not detected in the bile of any of the eel groups. In contrast, the 4-p-NP and the 1-OHPhe were found in all fish, whereas the other OH-PAHs and APs were not systematically detected. Because of the variability of the feeding status and the water level in the bile, some studies recommended the normalization of the biliary PAH metabolite concentration using the biliverdin or protein concentration in this biological matrix (Blahova et al., 2014; Nagel et al., 2012a; Ruddock et al., 2003). Investigating influences of the silvering process of the European eel on the PAH metabolite concentrations in bile, Nagel et al. (2012b) particularly recommended the use of a normalization. The protein concentration was chosen in this study to show levels normalized to protein concentrations as well as non-normalized measurements. The volume of bile of several eels was low and did not allow the analysis (YV: n = 6/10, YB: n = 8/10, YHI: n = 6/10 and SV: n = 6/15). Therefore, obtained results should be analyzed with caution. Nevertheless, they provide interesting data on the presence of contaminants poorly studied in the fish of the Loire estuary.

In the present study, the mean sums of OH-PAHs varied between 23.9 and 148.8 ng/mL bile with concentrations ranging from 25.5 to 100.5 ng/mL bile (11.3–20.4 µg/g prot.) for the 1-OHPyr and from 3.6 to 38.0 ng/mL bile (1.4–5.0 µg/g prot.) for the 1-OHPhe. European eels caught in the United Kingdom estuaries had 1-OHPyr and 1-OHPhe

<figure><img src="image_60.png" /><figcaption>**Fig. 5.** Profiles of the 12 PFAS in yellow European eels from three monitored sites in the Loire estuary (YV: Varades (n = 10); YB: Bellevue (n = 10); YHI: Haute Indre (n = 10)) and silver eels (SV; n = 15). PFAS proportions were calculated with the sum of 12 detected PFAS. Error bars correspond to standard deviation.</figcaption></figure>
contents from 120 to 7600 ng/mL bile and from 23 to 2096 ng/mL bile respectively (Ruddock et al., 2003). In eels from German rivers, the mean 1-OHPyr and 1-OHPhe detected values ranging from 10.9 to 2420.8 ng/mL bile and 52.7 to 632.4 ng/mL bile respectively (Nagel et al., 2012a). 1-OHNaph and 1-OHPyr were found in eels living in the Orbetello lagoon (Italy) at concentrations up to 8000 and 13,000 µg/mL bile, respectively (Gorbi et al., 2005). The contamination of eels in the Loire estuary is low compared with these levels depicted in eels from other European waters. These results are very surprising since the estuary of the Loire includes the second largest refinery of France, a coal-fired plant, an incinerator and a port with high activity (the 4th in France). All these activities are known to be sources of pollution, including PAHs, and the presence of metabolites in eel bile can provide information on the putative sources of PAH contamination. The 1-OHPyr, which is the prevalent compound detected in the eel bile in this study, is the main metabolite of pyrene. The pyrene is released by significant pyrolitic and petrogenic industrial processes. For this reason it is a common and widespread PAH (Ariese et al., 1993; Van Der Oost et al., 2003). The second greatest abundant metabolite found in bile eels from the Loire was the 1-OHPhe, one of the degradation metabolites of phenanthrene. This PAH is generated by the combustion of fossil fuels, in particular coal, oil and its refined products (O'Malley et al., 1996; Varanasi et al., 1989). Lower concentrations of 1-OHPhe compared with 1-OHPyr may be related to the fact that phenanthrene binds strongly to sediment particles reducing their bioavailability (Yuan et al., 2001). Furthermore, it was shown that some low molecular weight metabolites of phenanthrene were mainly excreted in the urine (Timbrell, 2000). Therefore, low levels in the bile may not reflect exactly the environmental exposure of the fish (Ruddock et al., 2003).

Among the three studied sites where yellow eels were caught, no differences between OH-PAH concentrations were found after the normalization of the protein contents of the bile (p > 0.05). Comparison between yellow and silver eels showed the same results, no difference was highlighted between the OH-PAH sums in bile (p > 0.05).

Among the five analyzed OH-PAHs, the 1-OHPyr was the most abundant (Fig. 6). It represented an average range between 36.0 and 80.8% of the sum of OH-PAHs. Although the 1-OHPyr is regarded as the most abundant compound present in fish bile and the best indicator of PAH

exposure in fish (Ariese et al., 1993; Beyer et al., 2010; Ruddock et al., 2003; Van Der Oost et al., 2003), this compound was not detectable in all eels in this study. Analysis of spatial influences showed differences for the 1-OHPyr, the 1-OHPhe and the 2-OHNaph. For the 1-OHPhe, YV eels had a significantly lower percentage of this metabolite than YB eels (p = 0.009). For the two other metabolites, it was the opposite; the YV eels had a significant higher proportion of 1-OHPyr and 2-OHNaph than the YB eels (p = 0.044 and p = 0.020) and also for this last than the YHI eels (p = 0.041).

Regarding life stage, differences were found for the 1-OHPhe and 2-OHNaph metabolites. SV eels had a greater percentage of 1-OHPhe than YHI and YV eels (p = 0.044 and p = 0.012). The 2-OHNaph was present in a lower proportion in the bile of SV eels than in YV eels (p = 0.017). This metabolite is an indicator of recent exposure to petrogenic compounds (Martínez-Gómez et al., 2013). Independent from the significant differences of individual OH-PAH proportions between sites and life stages, the distribution patterns of YV, YHI and SV eels seemed to be quite similar whereas those of YB eels was different, presenting high variability. These patterns seemed not to be balanced by the industrial, urban and agricultural activities. The origin of the OH-PAHs in eel bile and the organic compound behavior could be explained by the study of the PAHs in sediments.

APs and BPA were measured in the average range of 141.0–617.8 ng/mL bile and 36.8–3109.1 ng/mL bile respectively. Among the analyzed APs, two were mainly found: the 4-p-NP and the t-OP (concentrations ranging from 139.0 to 612.2 ng/mL bile and 2.8 to 15.1 ng/g bile, respectively). In the literature, no data seems to be available on wild eels. Nevertheless, some results were found for other fish species from the European waters. In the Bay of Biscay (Southwestern Europe), alkylphenolic compounds were measured in the bile of the thicklip grey mullet in the average range of 856–7483 and 21–153 ng/g of NP and OP levels respectively (Puy-Azurmendi et al., 2013). Biliary NP and t-OP were detected in *Alepocephalus rostratus* from the Western Mediterranean Basin with mean concentrations ranging from 42.5 to 67.1 ng/g bile for NP and from 7.9 to 14.2 ng/g bile for the t-OP (Koenig et al., 2013). Red mullets from polluted areas of the NW Mediterranean Sea showed mean concentrations that ranged from 0.34 to 28.3 µg NP/g bile and 30 to 250 ng t-OP/g bile (Martin-Skilton et al., 2006). The mean concentrations reported for red mullets caught at

<figure><img src="image_61.png" /><figcaption>**Fig. 6.** Profiles of the 5 OH-PAHs in yellow European eels from three monitored sites in the Loire estuary (YV: Varades (n = 6); YB: Bellevue (n = 8); YHI: Haute Indre (n = 6)) and silver eels (SV; n = 6). OH-PAH proportions were calculated considering the sum of 5 OH-PAHs. Error bars correspond to standard deviation.</figcaption></figure>
<figure><img src="image_62.png" /><figcaption>**Fig. 7.** Profile of the 5 APs in yellow European eels from three monitored sites in the Loire estuary (YV: Varades (n = 6); YB: Bellevue (n = 8); YHI: Haute Indre (n = 6)) and silver eels (SV; n = 6). AP proportions were calculated with the sum of 5 APs. Error bars correspond to standard deviation.</figcaption></figure>

several sites along the Spanish Mediterranean coast were 358.4–923.3 ng/g for the NP and 18.8–63.1 ng/g bile for t-OP (Martínez-Gómez et al., 2013).

Compared with these data, eels from the Loire estuary were less exposed to APs than other fish caught in different waters from Europe. Regarding BPA, it was found at concentrations similar to those found in some other studies. For example, in the bile from male breams collected in Netherlands, levels of BPA measured were 32.4 to 3082.3 ng/mL bile (Houtman et al., 2004).

Among the three sites of the Loire estuary, BPA was not detected in the bile of the YHI eels. The YV eels had high levels of BPA, 10 fold higher than the YB eels (p = 0.007). Regarding the life stage differences, the YV eels presented high BPA levels in the bile (p < 0.001), around 200 fold greater than the SV eels. Once again, these results are difficult to correlate to the different natures of site activities.

For the AP sums measured in the bile, no difference was found between the yellow eels. The SV eels had once again a lower mean AP sum, significantly different than those of the YHI eels (p = 0.022).

The AP pattern is presented in Fig. 7. The more abundant AP in the eel bile was the 4-p-NP with ranges from 70.2 to 99.2% followed by the 4-tOP with a percentage comprised between 0.5 and 29.8%. As with the mean AP concentrations, no variation was found between profiles of the yellow eels from the three sampling sites. Regarding the life stage influence, the SV eels presented no significant difference with the YV eels and different patterns compared with the two other yellow groups. The SV eels had a greater percentage of 4-tOP than the YHI and YB eels (p < 0.002) and a lower percentage of the 4-pNP (p < 0.005). For APs, and independent from the significant differences of individual proportions between sites and life stages, the SV eels showed patterns slightly different than the yellow eels with high

<figure><img src="image_63.png" /><figcaption>**Fig. 8.** Principal component analysis (PCA) of biometric parameters, contaminants (dl- and ndl-PCBs, PBDEs and PFAS expressed in ng/g dw) in muscle and contaminants (BPA, OH-PAHs and APs expressed in µg/g prot.) in bile of yellow eels from 3 sampling sites in the Loire estuary (Varades (n = 10); Bellevue (n = 10); Haute Indre (n = 10)) and of silver eels (n = 15). A: correlation loadings (BW: body weight; TL: total length; LW: lipid weight). B: sample representation (circles = yellow eels from Varades; squares = yellow eels from Bellevue; triangles = yellow eels from Haute Indre; diamonds = silver eels).</figcaption></figure>
variability. The difference as well as the variability could be explained by the fact that migrating silver eels could come from upstream areas more or less polluted. The sediment analysis could confirm the similar AP ex- position of the eels in the different sites.

### 3.4. PCA results

The correlation loading and sample representation are shown in Fig. 8 (respectively panels A and B). The first two principal components (respectively PC1 and PC2) described 48% of the total variability among the eels. PC1 and PC2 represented respectively 29% and 19%. The corre- lation loadings (Fig. 8A) highlighted that the first PC was positively correlated to TL, BW, LW and PBDE. The second PC was positively corre- lated to PFAS and negatively to GSI. APs, OH-PAHs, BPA and PI presented low weight in the PC construction. Biometric parameters BW and TL cor- related to each other and to ndl-, dl-PCBs and PFAS. All these variables were negatively correlated to HSI. Concerning LW, it appeared to be positively correlated to both GSI and PBDEs. Regarding the sample pre- sentation (in Fig. 8B), the eels were clustered according to the life stages. About the sampling sites of the yellow eels, the eels from Varades were separated from the eels from Bellevue whereas the eels from Haute Indre were dispersed among the others. The comparison of Fig. 8A and Fig. 8B underlines that the eels from Bellevue were the biggest and the most highly contaminated by PFAS, ndl- and dl-PCBs (dw basis). Silver eels were highly contaminated by PBDEs with the highest lipid contents and GSI. The trends for the yellow eels from Varades and Haute Indre were contrary to the silver eels, e.g. low levels of PBDEs, LW and GSI. These observations are consistent with result explanations depicted in the previous paragraphs.

### 3.5. Human health risk assessment

Levels of 6 PCBi and TEQ values were showed in Table 3. Commission Regulation (EC) No 1259/2011 of 2 December 2011 had set maximum levels for dioxins and dioxin-like PCBs in a range of foodstuffs including wild eels (European Commission, 2011). In this study, the PCDD/F levels were not determined in the eels from the estuary, thus only dl-PCB TEQs were estimated. However, Total TEQ (1998) and dl-PCB TEQ (1998) values were determined using prediction models for fish (log TEQ DL-PCBs = −1.167(±0.051) + 0.932(±0.022) × logΣPCBi + PCB118) and log TEQ Total = −1.128(±0.053) + 0.929(±0.023) × logΣ(PCBi + PCB118) (Babut et al., 2009).

The maximum levels set by the European Commission for specific oil species, including wild eels, were 3.5 pg WHO2005 TEQ/g ww for the PCDD/F TEQ and 10 pg WHO2005 TEQ/g ww PCDD/F TEQ + dl-PCBs TEQ (Total TEQ). Because no dl-PCB TEQ maximum value was set by the European Commission, the results were compared with the Total TEQ. Among the yellow eels, 7 individuals (4 YB and 3 YHI) had dl-PCB TEQs (2005) higher than the Total TEQ (2005) value (ranging from 10.85 to 16.93 pg WHO2005 TEQ/g ww) with a mean of 8.06 ± 4.67 pg WHO2005 TEQ/g ww for the YB and 7.18 ± 5.58 pg WHO2005 TEQ/g ww for the YHI. The YV eels had the lowest dl-PCB TEQ values (mean of 2.07 ± 1.21 pg WHO2005 TEQ/g ww). The SV eels had a dl-PCB TEQ mean (8.45 ± 2.05 pg WHO2005 TEQ/g ww) close to the

maximum value. Nevertheless, 2 individuals showed concentrations above this regulation (10.04 and 14.55 pg WHO2005 TEQ/g ww).

Few differences were found between calculated and predicted dl-PCB TEQs 1998, even if the prediction seems to be more accurate for the yellow eel values. Regarding the predicted Total TEQ (98), for the yellow and silver eels, the means estimated were above the 10 pg WHO1998 TEQ/g ww Total TEQ.

The maximum level for the sum of the 6 PCB indicators (PCB 28, 52, 101, 138, 153 and 180) was established at 300 ng/g ww (European Commission, 2011). One third of the yellow eels had a concentration of 6 PCBi higher than this value (ranging from 318.00 to 746.16 ng/g ww). The YB and YHI eels had the highest levels (mean of 329 ± 208 and 256 ± 170 ng/g ww, respectively) with 6/10 and 4/10 of the eels respectively, presenting concentra- tions higher than the maximum level. YV eels were the least con- taminated (mean of 63 ± 28 ng/g ww). Silver eels presented intermediate levels and all individuals presented values lower than the maximum level.

The results indicated a potential exposure to PCBs through eel con- sumption, especially with the yellow eels. The tolerable daily intake (TDI) of 10 ng PCBi/kg body weight/day set by the French Food Safety Agency represents an intake of 700 ng/day for a 70 kg person or 150 ng/day for a child of 15 kg (French Food Safety Agency, 2010). Blanchet-Letrouvé et al. (2014) recommended the consumption of a portion of 150 g per month which would correspond to a daily intake of 1004 ng/day considering the mean concentration of all the eels (ranging from 109.0to 3610.5 ng/day). In this study, it could then be recommended to limit the consumption to one portion of 105 g per month. For the sensitive populations (a portion of 60 g every two months), the average dietary intake would be 201 ng/day which was above the specific recommendation of the agency. The portion should be restricted to 45 g every two months. Considering the local practices (mean annual frequency of 2.6 times per year; Blanchet-Letrouvé et al., 2014) and the results of this study, it is possible to calculate an es- timated dietary daily intake of PCBi of 222 ng/day. As reported by Blanchet-Letrouvé et al. (2014), considering the local practices, the risk seems to be moderate for adults but exists for sensitive populations. Nevertheless, according to the TDI, the recommendations of the agency about the consumed portions are not really adapted to the eels fished in the Loire estuary and should be revised.

The assessment of consumer risks from the intake of PFAS was also realized. The EFSA suggested that dietary exposure was the main expo- sure route to PFAS in the human population and that fish and other sea- food were considered as the major contributors (European Food Safety Authority, 2012). In 2008, a risk assessment for the two major PFAS, the PFOS and the PFOA was performed by the EFSA Scientific Panel on Con- taminants in the Food Chain (CONTAM Panel). The authority concluded that it was not probable that adverse effects of PFOS and PFOA are occur- ring in the general population. However to assess potential impacts on human health, the CONTAM Panel suggested a tolerable daily intake (TDI) of 150 ng/kg bow per day for PFOS and 1500 ng/kg bw per day for PFOA (Authority, 2008). These values were far above the concentra- tions found in the eels from the estuary. On average, fish consumption per capita in Europe was 22.1 kg (Statistics and Information Service of the Fisheries and Aquaculture Department, 2010). Thus, for a 70 kg

<table><caption>**Table 3**
Mean concentrations of 6 indicator PCBs (Σ6 PCBi), dioxin-like PCB TEQ 1998 (dl-PCB TEQ (98)), dioxin-like PCB TEQ 2005 (dl-PCB TEQ (05)) and predicted TEQ concentrations for dioxin-like PCBs (predicted dl-PCB TEQ (98)) and Total TEQ (predicted Total TEQ (98)) according to the life stage.</caption><thead><tr><th></th><th>Σ<sub>6</sub> PCBi<br>(ng/g ww)</th><th>dl-PCBs TEQ (98)<br>(pg TEQ WHO 1998/g)</th><th>dl-PCBs TEQ (05)<br>(pg TEQ WHO 2005/g)</th><th>Predicted dl-PCB TEQ (98)<br>(pg TEQ WHO 1998/g)</th><th>Predicted Total TEQ (98)<br>(pg TEQ WHO 1998/g)</th></tr></thead><tbody><tr><td>Yellow eels<br>(n = 30)</td><td>216.34 ± 188.80<br>(22.53–746.16)</td><td>9.37 ± 7.61<br>(1.04–25.86)</td><td>5.77 ± 4.91<br>(0.50–16.93)</td><td>10.85 ± 8.80<br>(1.47–33.97)</td><td>11.67 ± 9.44<br>(1.59–36.43)</td></tr><tr><td>Silver eels<br>(n = 15)</td><td>189.97 ± 34.68<br>(144.14–253.54)</td><td>12.35 ± 2.50<br>(9.26–18.99)</td><td>8.45 ± 2.05<br>(6.00–14.55)</td><td>9.91 ± 1.68<br>(7.71–13.05)</td><td>10.67 ± 1.80<br>(8.31–14.04)</td></tr></tbody></table>

Mean ± standard deviation; Σ6 PCBi (PCB 28, 52, 101, 138, 153, 180); min-max values are presented into brackets.
person, with an average dietary daily intake of 60.5 g of fish, the potential dose of exposures of PFOS represented 10.4% and 14.4% of the TDI for consumption of yellow and silver eels respectively. For the PFOA, the potential dose exposures represented 0.004% and 0.012% of the TDI for the consumption of the yellow and silver eels respectively.

Regarding all the recommendations and the contaminant levels in eel muscles, a total of 13 individuals out of 45 were at least above one legal consumption limit and could be considered as unfit for human consumption.

## 4. Conclusion

The Loire, being considered as a wild river not highly polluted in France, has long been little studied. This study presented several data on biological status and monitored contaminant levels on the eels from the Loire estuary, associating risks to the fish consumers. It constitutes an important contaminant screening realized on wild European eels in a French river. Economic issues, particularly in terms of water quality, but also quality of the consumer products, such as fish, have led local and national public authorities to run programs for monitoring the estuary. It is another step in the monitoring of the environmental quality of the Loire estuary, which began with an initial study by Blanchet-Letrouvé et al. (2014). It would be interesting to follow the evolution over time of the levels and profiles of these POPs to check the progress thereof and thus, characterize potential sources of contamination. The results of this study showed that the levels of PCBs, PBDEs and PFAS in eel muscles from the Loire estuary were intermediate compared with eels from other European waters. The OH-PAH and AP concentrations in the eel biles could be considered as low compared with the European data whereas the BPA showed similar levels. The eels presented important variability of contaminant levels whatever the monitored location leading to attenuated differences between the sites and the life stages. The majority of eels exceeded the PFOS and PBDE EQS biota and a medium risk for the consumer has been demonstrated with 20% of the eels presenting values higher than the TEQ maximum level (33% considering the yellow eels only). A limited consumption of eels should be advised. To complete this first study, monitoring of sediment and water quality in the estuary should be undertaken.

## Acknowledgments

The authors would like to express their special thanks to the region Pays de la Loire, the SMIDAP (Syndicat Mixte pour le Développement de l'Aquaculture et de la Pêche en Pays de la Loire, arrêté N° 2012-06091) and the AADPPMFEDLA (Association Agréée Départementale des Pêcheurs Professionnels Maritimes et Fluviaux en Eau Douce de Loire-Atlantique) for their technical and financial support. The technical assistance of Claire Favreau and Emilie Perrat is gratefully acknowledged. The authors thank Darci (McCardle) Couderc for the English language revision of the manuscript.

## Appendix A. Supplementary data

Supplementary data to this article can be found online at http://dx.doi.org/10.1016/j.scitotenv.2014.09.053.

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<figure><img src="image_64.png" /></figure>
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Review

# Is there any consistency between the microplastics found in the field and those used in laboratory experiments? ☆

<figure><img src="image_66.png" /></figure>

CrossMark

Nam Ngoc Phuong <sup>a, b</sup>, Aurore Zalouk-Vergnoux <sup>a, b, *</sup>, Laurence Poirier <sup>a, b</sup>, Abderrahmane Kamari <sup>a, b</sup>, Amélie Châtel <sup>a, b</sup>, Catherine Mouneyrac <sup>a, b</sup>, Fabienne Lagarde <sup>c</sup>

ARTICLE INFO

Article history:

Received 12 October 2015

Received in revised form

16 December 2015

Accepted 17 December 2015

Available online xxx

Keywords:

Microplastics

Field samples

Laboratory exposures

Ingestion

Biological effects

A B S T R A C T

The ubiquitous presence and persistency of microplastics (MPs) in aquatic environments are of particular concern since they represent an increasing threat to marine organisms and ecosystems. Great differences of concentrations and/or quantities in field samples have been observed depending on geographical location around the world. The main types reported have been polyethylene, polypropylene, and poly- styrene. The presence of MPs in marine wildlife has been shown in many studies focusing on ingestion and accumulation in different tissues, whereas studies of the biological effects of MPs in the field are scarce. If the nature and abundance/concentrations of MPs have not been systematically determined in field samples, this is due to the fact that the identification of MPs from environmental samples requires mastery and execution of several steps and techniques. For this reason and due to differences in sampling techniques and sample preparation, it remains difficult to compare the published studies.

Most laboratory experiments have been performed with MP concentrations of a higher order of magnitude than those found in the field. Consequently, the ingestion and associated effects observed in exposed organisms have corresponded to great contaminant stress, which does not mimic the natural environment. Medium contaminations are produced with only one type of polymer of a precise sizes and homogenous shape whereas the MPs present in the field are known to be a mix of many types, sizes and shapes of plastic. Moreover, MPs originating in marine environments can be colonized by organisms and constitute the sorption support for many organic compounds present in environment that are not easily reproducible in laboratory. Determination of the mechanical and chemical effects of MPs on organisms is still a challenging area of research. Among the potential chemical effects it is necessary to differentiate those related to polymer properties from those due to the sorption/desorption of organic compounds.

© 2015 Elsevier Ltd. All rights reserved.

1. Introduction

Plastics is a generic name encompassing most of the synthetic organic polymers exhibiting the property of plasticity. These products present many advantages: they are inexpensive, water and corrosion-resistant, chemically inert, easily molded, and they exhibit good thermal and electrical insulating properties. All these features explain why they are massively used in our daily lives. Since the mid-20th century, several million tons of plastics have

been produced (Thompson et al., 2009) and in 2013, worldwide plastic production was estimated at 288 million tons (Free et al., 2014). Whilst the societal benefits of plastic are undeniably far- reaching (Andrady and Neal, 2009), this valuable commodity is nonetheless the subject of increasing environmental concern. Indeed, plastics present many disadvantages: being nonrenewable resources and sources of contamination by additive compounds; undergoing embrittlement at low temperatures and deformation under loads; going through a costly recycling process; being highly resistant to degradation, etc. It has been estimated that ten percent of the plastics produced end up in the ocean. Jambeck et al. (2015) recently made the approximate calculation that in 2010 alone, 4.8 to 12.7 million metric tons of plastic wastes entered the ocean and pointed out that a steady increase is to be expected in the coming

<footnote><sup>a</sup> Laboratoire de Mer, Molécules, Santé (MMS, EA 2160), Université de Nantes, Nantes F-44322, France</footnote>

<footnote><sup>b</sup> Université Catholique de l'Ouest, Angers F-49000, France</footnote>

<footnote><sup>c</sup> Institut des Molécules et Matériaux du Mans (IMMM, UMR CNRS 6283), Université du Maine, Avenu Olivier Messiaen, Le Mans F-72085, France</footnote>

<footnote>* This paper has been recommended for acceptance by Klaus Kummerer.</footnote>

<footnote>* Corresponding author.
E-mail address: aurore.zalouk-vergnoux@univ-nantes.fr (A. Zalouk-Vergnoux).</footnote>
years. After less than a century of existence, plastic debris already represent from 60 to 80% of marine litter depending on the locations investigated (Derraik, 2002). Once in the environment, macrodebris undergo mechanical (erosion, abrasion), chemical (photoroxidation, temperature, corrosion) and biological (degradation by microorganisms) actions (Andrady, 2011; Costa et al., 2010; Zettler et al., 2013). These different degradation processes lead to their fragmentation into microplastics (MPs), which accumulate in the environment.

The term of MP was initially suggested by Thompson (Thompson et al., 2004). The size defining MPs varies according to authors with diameters of >1.6 µm (Ng and Obbard, 2006) and <1 mm (Browne et al., 2007, 2010; Claessens et al., 2011), <2 mm (Ryan et al., 2009), 2–6 mm (Derraik, 2002), <5 mm (Barnes et al., 2009), <10 mm (Graham and Thompson, 2009). Nowadays, most researchers agree with the definition of MPs proposed by Arthur et al. (2009) of MPs as particles in a size range of less than 5 mm. Those originating from the fragmentation of larger plastic items are defined as secondary MPs whereas primary MPs include all microsized particles entering the environment: such as fibers, industrial pellets and microbeads from cosmetics, for example (Andrady, 2011).

Whatever the definition of their size, MPs represent a very broad range of polymers. The most commonly used plastic materials are polyethylene (PE), PP, polyvinyl chloride (PVC), PS and polyethylene terephthalate (PET). They represent approximately 90% of the total world production (Andrady and Neal, 2009) and, as non-biodegradable polymers, they are expected to be among the most widely represented in sampled MPs. What happens to them in the environment will probably differ according to their chemical nature and physical properties. Being buoyant in water; PE and PP, float in seawater, mainly affect ocean surface and deposit ashore (Engler, 2012; Thompson et al., 2009) whereas PVC, which is denser than seawater, affects seabed, often next to the source (Engler, 2012). According to the most recent ten years of research, it appears that all natural habitats from pole to pole are affected by the presence of MPs (Wright et al., 2013).

The ubiquitous presence and persistency of MPs in aquatic environments are of particular concern since they represent an increasing threat to marine organisms and ecosystems. To evaluate this threat and the potential impacts of MPs on aquatic organisms, a number of laboratory experiments have been performed over recent years to mimic exposure (Nobre et al., 2015; Van Cauwenberghe et al., 2015). The global aim of this review is to compare the available literature data on MPs (size, form, quantity or concentration) having been sampled around the world with those having been employed in laboratory experiments. In this paper, we try to answer the following questions: i) Which MPs are found in the different environmental compartments (water, sediment, biota) of the marine ecosystems? ii) What are the laboratory conditions of exposure and putative ingestion of MPs by organisms, as well as their biological transfer and trophic transfer? iii) Are experimental laboratory exposures indeed consistent with this exposome? iv) Do the toxicity effects reported on a wide range of aquatic organisms actually reflect environmental reality?

## 2. Which MPs are found in the different environmental compartments?

### 2.1. MPs in marine waters

Due to their small size and capacity to float on the surface of seas and oceans, MPs are distributed in all marine ecosystems. At least 29 marine areas in the world have been investigated (detailed in Table 1).

Differences between sites are observed in MP concentrations and/or quantities. The Pacific Ocean is the most widely sampled area. When comparable, the measured concentrations in this ocean vary from almost 27,000 (Eriksen et al., 2013) to 448,000 (Goldstein et al., 2013) particles per km² and from 0.004 (Doyle et al., 2011) to 9200 (Desforges et al., 2014) particles per m³. This uneven distribution is partially accounted for by the well-known zone of concentration in the North Pacific Central Gyre, where the large-scale presence of plastic debris has previously been highlighted. Indeed, several studies have shown that quantities and distributions of MPs are pronouncedly dependent on geophysical processes (wave, wind, water current; Goldstein et al., 2013; Wright et al., 2013). For example, the highest densities of MPs were detected in the Northeast Pacific Ocean under low-wind conditions (Goldstein et al., 2013). In the Atlantic Ocean, all reported plastic concentrations have been significantly lower than in Pacific areas with only 1500 particles per km² (Law et al., 2010) and 2.5 particles per m³ (Lusher et al., 2014) in the most polluted zones. The Mediterranean Sea is another high spot of MP presence and a recent publication reported high levels of concentrations compared to the Atlantic Ocean (62,000 particles per km² for the lowest concentration; Collignon et al., 2014). Considerable research activity in the European zone has recently taken place since the Mediterranean Sea is a "closed sea" surrounded by urbanized areas, which may explain the high quantity of MPs (Cózar et al., 2015).

A reason other than geographic disparities for which it remains difficult to directly compare the reported MP concentrations/quantities is that the sampling techniques and analytical methodologies used in sample analysis are highly varied. For water samples, differences in sampling protocols lead to obvious difficulties of comparison between studies. Indeed, concentration data are usually expressed in terms of number of particles either by volume of filtered water (cubic meters) in the case of water pumping; or by area of waters covered (square meters) when using a trawl along a transect. The unit difference underlines the lack of standardized sampling methodology. Number of particles per km² implies that samples came mostly from water surfaces whereas particles per m³ are likely to correspond to water column sampling. Given the wide range of polymer densities, these differences in sampling depth will surely affect results in terms of both concentration and polymer nature, even though Carson et al. (2011) have shown that in the water column, 95% of small plastic debris were concentrated in the top 15 cm. For example, the use of trawl enables MPs to be sampled on large surface areas. On the other hand, while pumped samples are more representative of MP distribution in the water column, this is only the case at a sampling point. Moreover, as seen in a few recent studies, depending on the protocol used the mesh aperture of nets can vary from around 330 µm (the most commonly used so far) to less than 50 µm (Hidalgo-Ruiz et al., 2012). As high volumes of sieved water lead to a rapid clog the lowest sizes of mesh aperture are less frequently employed. In addition, extraction and analysis of particles smaller than 330 µm is more challenging (Desforges et al., 2014).

Despite differences of MP concentrations/quantities, and even though the MP type was mentioned in only 6 studies out of 25, differences in MP types can be noticed are apparent, since PE and PP fragments have been found in sea water samples from America and Australia, while PE and PS have been observed in Europe.

### 2.2. MPs in sediment/sand samples

It is well-known that sediments are the long-term ultimate sink for of contaminants (Chapman and Wang, 2001). Table 2 illustrates some examples of MPs reported in the sediment compartment over the last 40 years.

As it is shown in Table 2, America, Asia and Europe are the most
<table><caption>**Table 1**
Sampling continent and area, quantity or concentration, type or form of MPs found in sea water samples and references of the corresponding studies.</caption><thead><tr><th>Continent</th><th>Area</th><th>Quantity or concentration</th><th>Microplastics (type or form)</th><th>References</th></tr></thead><tbody><tr><td rowspan="14">America</td><td>Western North Atlantic</td><td>3500 particles/km²</td><td>Pellets 2.5–5 mm</td><td>Carpenter and Smith (1972)</td></tr><tr><td>North Pacific Ocean</td><td>0.004–0.19 particles/m³</td><td>Fragment predominant</td><td>Doyle et al. (2011)</td></tr><tr><td>Atlantic</td><td>&lt;0.1 particles/m³</td><td>Fragment predominant</td><td></td></tr><tr><td>South Pacific SG</td><td>26,898 pieces/km²</td><td>Fragment predominant</td><td>Eriksen et al. (2013)</td></tr><tr><td>North Pacific SG</td><td>32.76 particles/m³</td><td>Fragment predominant</td><td>Goldstein et al. (2012)</td></tr><tr><td>Equatorial Atlantic Ocean</td><td>0.01 particles/m³</td><td>Fragment predominant</td><td>Ivar do Sul et al. (2013)</td></tr><tr><td>Southern Californian</td><td>3.92 items/m³</td><td>Fragment predominant</td><td>Lattin et al. (2004)</td></tr><tr><td>North Atlantic Ocean</td><td>1534 pieces/km²</td><td>PE + PP 99%</td><td>Law et al. (2010)</td></tr><tr><td>Caribbean Sea</td><td>1414 pieces/km²</td><td>PE + PP 99%</td><td></td></tr><tr><td>North Pacific Gyre</td><td>2.23 particles/m³</td><td>PP monofilament predominant</td><td>Moore et al. (2001)</td></tr><tr><td>North Pacific Central Gyre</td><td>334,271 fragments/km²</td><td>Fragment 58.5%</td><td></td></tr><tr><td>Southern California</td><td>7.25 particles/m³</td><td>Fragment 92.7–100%</td><td>Moore et al. (2002)</td></tr><tr><td>North Pacific offshore</td><td>0.43–2.23 particles/m³</td><td>Fragment predominant</td><td>Moore et al. (2005)</td></tr><tr><td>North Pacific, inshore</td><td>5–7.25 particles/m³</td><td></td><td></td></tr><tr><td>Pacific Ocean</td><td>370,000 particles/km²</td><td>Fragment 92.6%</td><td>Shaw and Day (1994)</td></tr><tr><td>NE Pacific Ocean</td><td>8 to 9200 particles/m³</td><td>Fibers 75%</td><td>Desforges et al. (2014)</td></tr><tr><td>Northeast Pacific Ocean</td><td>0.021–0.448 particles/m²</td><td></td><td>Goldstein et al. (2013)</td></tr><tr><td rowspan="2">Asia</td><td>Western Pacific Ocean</td><td>87,000 pieces/km²</td><td>Fragment 56%</td><td>Yamashita and Tanimura (2007)</td></tr><tr><td>China</td><td>0.167 particles/m³</td><td>Fibers, granules</td><td>Zhao et al. (2014)</td></tr><tr><td>Australia</td><td>Australia</td><td>4256–8966 pieces/km²</td><td>PE + PP 98.5%</td><td>Reisser et al. (2013)</td></tr><tr><td rowspan="10">Europe</td><td>North Sea</td><td>50–100 fibers/l</td><td>Fibers predominant</td><td>Buchanan (1971)</td></tr><tr><td>Mediterranean Sea</td><td>0.116 particles/m²</td><td>Filaments, PS</td><td>Collignon et al. (2012)</td></tr><tr><td>Italian coast</td><td>0.62 particles/m³</td><td>Plastic fragment</td><td>Fossi et al. (2012)</td></tr><tr><td>Plymouth, UK</td><td>&lt;0.04 pieces/m³</td><td>Plastic fibers</td><td>Thompson et al. (2004)</td></tr><tr><td>French-Belgian-Dutch</td><td>0.1–0.7 particles/l</td><td>LDPE, HDPE and PS predominant</td><td>Van Cauwenberghie et al. (2015)</td></tr><tr><td>Mediterranean–Corsica</td><td>6.2 particles/100 m²</td><td>Filaments, PS</td><td>Collignon et al. (2014)</td></tr><tr><td>Northeast Atlantic Ocean</td><td>2.46 particles/m³</td><td>Plastic fibers &gt; 80.3%</td><td>Lusher et al. (2014)</td></tr><tr><td>Mediterranean Sea</td><td>243,853 items/km²</td><td>Fragment 87.7%</td><td>Cozar et al. (2015)</td></tr><tr><td>Western Mediterranean</td><td>0.15 items/m³</td><td>Foam, filament, pellet</td><td>Lucia et al. (2014)</td></tr></tbody></table>

<table><caption>**Table 2**
Sampling continent and area, quantity or concentration, type or form of MPs found in sediment/sand samples and references of the corresponding studies.</caption><thead><tr><th>Continent</th><th>Area</th><th>Quantity or concentration</th><th>Microplastic (type or form)</th><th>References</th></tr></thead><tbody><tr><td rowspan="10">America</td><td>Kamilo Beach**</td><td>248 items/m³</td><td>PE 85%, PP 14%</td><td>Carson et al. (2011)</td></tr><tr><td>Florida**</td><td>116–214 piece/l</td><td>Fragment predominant</td><td>Graham and Thompson (2009)</td></tr><tr><td>Maine**</td><td>105 piece/l</td><td>Fragment predominant</td><td></td></tr><tr><td>Canada*</td><td>&lt;10 particles/m</td><td>PE predominant, virgin pellets</td><td>Gregory (1983)</td></tr><tr><td>Bermuda*</td><td>&gt;5000 particles/m</td><td>PE predominant, virgin pellets</td><td></td></tr><tr><td>Pacific Ocean*</td><td>27 items/m² (1–4.75 mm)</td><td>Fragment 89%</td><td>Hidalgo-Ruiz and Thiel (2013)</td></tr><tr><td>Noronha Archipelago*</td><td>15 particles/kg</td><td>Fragment 65%; pellet 23%</td><td>Ivar Do Sul et al. (2009)</td></tr><tr><td>Canada**</td><td>2–8 pieces/g</td><td>Plastic fibers</td><td>Mathalon and Hill (2014)</td></tr><tr><td>Hawaiian beaches*</td><td>43.4 particles/l</td><td>fragment 87%, plastic pellets 11%</td><td>McDermid and McMullen (2004)</td></tr><tr><td rowspan="12">Asia</td><td>Oman Gulf*</td><td>&gt;100 particles/m²</td><td>PE predominant, virgin pellets</td><td>Khordagui and Abu-Hilal (1994)</td></tr><tr><td>Arabian Gulf*</td><td>To 80,000 particles/m²</td><td>PE predominant, virgin pellets</td><td></td></tr><tr><td>Sea of Japan*</td><td>8–17 particles/m²</td><td>Fragment 41%</td><td>Kusui and Noda (2003)</td></tr><tr><td>Singapore**</td><td>0–4 particles/sample</td><td>PE and PS predominant</td><td>Ng and Obbard (2006)</td></tr><tr><td>Indian Ocean**</td><td>81.43 mg/kg</td><td>Fragment 100%</td><td>Reddy et al. (2006)</td></tr><tr><td>India*</td><td>68.83 items/m²</td><td>Fragment predominant</td><td>Jayasiri et al. (2013)</td></tr><tr><td>Singapore**</td><td>36.8 particles/kg</td><td>PE, PP, nylon and PVC</td><td>Mohamed Nor and Obbard (2014)</td></tr><tr><td>China*</td><td>4137.3 particles/m³</td><td>Fibers, granules</td><td>Zhao et al. (2014)</td></tr><tr><td>Korea*</td><td>Up to 27,606 particles/m²</td><td>PS expanded &gt; 96%</td><td>Lee et al. (2013)</td></tr><tr><td>SW Indian Ocean**</td><td>26 particles/l</td><td>Plastic fibers</td><td>Woodall et al. (2014)</td></tr><tr><td>Australia</td><td>New Zealand*</td><td>&gt;1000 particles/m</td><td>PE and PP predominant</td><td>Gregory (1977)</td></tr><tr><td rowspan="10">Europe</td><td>Tamar Estuary, UK**</td><td>32 items/sample</td><td>&lt;1 mm, PVC 26%; Polyester 35%</td><td>Browne et al. (2010)</td></tr><tr><td>Belgian coast**</td><td>&lt;391 particles/kg</td><td>&lt;1 mm, fiber 59%, granule 25%</td><td>Claessens et al. (2011)</td></tr><tr><td>Russian beaches*</td><td>5–10 particles/m²</td><td>Fragment 55.6%</td><td>Kusui and Noda (2003)</td></tr><tr><td>Plymouth, UK**</td><td>&gt;10 pieces/l</td><td>Plastic fibers</td><td>Thompson et al. (2004)</td></tr><tr><td>Maltese coast*</td><td>&gt;1000 particles/m²</td><td>PE, plastic pellets</td><td>Turner and Holmes (2011)</td></tr><tr><td>French–Belgian–Dutch**</td><td>0.3–11.7 particles/kg</td><td>LDPE, HDPE and PS predominant</td><td>Van Cauwenberghie et al. (2015)</td></tr><tr><td>Venice, Italy**</td><td>672–2175 pieces/kg</td><td>PE + PP 82%</td><td>Vianello et al. (2013)</td></tr><tr><td>Slovenia*</td><td>Up to 155.6 particles/kg</td><td>MPs (1–5 mm) &gt; 74%; fibers, fragments</td><td>Laglbauer et al. (2014)</td></tr><tr><td>North Sea**</td><td>1.3–2.3 particles/kg</td><td>PP, PE, PET, PVC, PS and polyamide</td><td>Dekiff et al. (2014)</td></tr><tr><td>NE Atlantic Ocean**</td><td>324 particles/l</td><td>Plastic fibers (polyester 53.4%)</td><td>Woodall et al. (2014)</td></tr><tr><td>Mediterranean Sea**</td><td>350 particles/l</td><td></td><td></td></tr></tbody></table>

widely studied continents as regards the quantification of MPs in sandy samples and sediments. Studies in Europe deal mainly with sediments (8 out of 11) whereas studies on sandy samples are

predominant in Asia (6 out of 10) and in America there are as many studies on sandy samples as on sediments. Concerning the geographical comparison of MP quantity in sandy/sediment

<footnote>*: sandy samples (mineral deposit); **: sediment samples (deposit rich in organic matter).</footnote>
samples, the location of studied areas seemed to be the main factor influencing MP distribution in the field. When comparable, the measured concentrations of MPs in the sediments ranged widely, from almost 0.3 (Van Cauwenberghe et al., 2015) to 2175 (Vianello et al., 2013) particles per kg and from 27 (Hidalgo-Ruiz and Thiel, 2013) to 80,000 (Khordagui and Abu-Hilal, 1994) particles per m² in sandy samples.

As for water samples, due to differences in sampling techniques it remains difficult to directly compare the reported MP concentrations/quantities or types. Sandy/sediment samples are usually collected from the coast at low tide (Hidalgo-Ruiz et al., 2012). The depth at which they are found, the masses of the samples and the tide level have tended to vary considerably between the different studies. For sampling at low tide from the coast, different types of tools are usually used according to sample location: stainless-steel spoons, spatula and box corer. For sampling at high tide, it is possible to use a corer and bottom trawl. The diversified sampling methods lead to results expressed in different units, making it difficult to compare between published studies. Quantification of MPs in sandy samples is usually expressed in number of particles per unit of distance and surface (Gregory, 1983; Hidalgo-Ruiz and Thiel, 2013; Kusui and Noda, 2003) while for sediment samples, MP quantities are usually expressed in number of particles per unit of volume (Graham and Thompson, 2009; Woodall et al. 2014) or per m³ (Carson et al. 2011). However, number of particles per unit of mass has been used for both types of samples, i.e. Ivar Do Sul et al. (2009) for sandy samples and Claessens et al. (2011) as well as Vianello et al. (2013) for sediments. This discrepancy is once again due to a lack of standardization for sampling protocols.

In addition to differences between sample collections, sediment sample preparation can lead to a variability of results between studies. Filtration and/or extraction protocols based on separation by density are usually employed for the quantification of MPs from sand and sediment samples. However, different sizes of sieves and filters have been reported. For example, Claessens et al. (2011) studied MPs between 38 µm and 1 mm and Laglbauer et al. (2014) from 1 to 5 mm. Besides, solutions of water saturated with sodium chloride (Thompson et al., 2004) or sodium iodide (Claessens et al., 2013) have been used to extract MPs from sediment samples. The use of these two different salts generates different densities of solutions (1.2 g cm⁻³ for saturated sodium chloride and 1.6 g cm⁻³ for sodium iodide at 3.3 M) leading to variations in MP extraction capacity.

Concerning the different MP types found in sand/sediment samples, when mentioned in studies, PE has been largely predominant. The presence of PE is quite surprising since it is a low-density polymer and should float at the sea surface. This observation supports the hypothesis of a possible colonization of organisms on these MPs leading to increased density (Zettler et al., 2013; Harrison et al., 2014).

## 2.3. MPs in marine organisms

The MPs found in marine organisms may be the result of their ubiquitous presence in highly abundant marine ecosystems. By using a manta trawl lined with 333 µm mesh at 11 random sites of the North Pacific Ocean, Moore et al. (2001) showed that although plankton organisms were approximately five times more abundant than plastic particles, the total mass of plastic was six times greater than that of plankton. These results underlined the fact that MPs have a high potential to affect organisms due to their numerous contacts with one another, as well as with the different organisms.

The presence of MPs in organisms from marine ecosystems is now well-established (Desforges et al., 2015; Mathalon and Hill, 2014; Van Cauwenberghe and Janssen, 2014). A majority of

studies have focused on the ingestion of MPs in various species ranging from invertebrate to marine mammal. However, to date, data on the biological effects of MPs on marine organisms are still scarce. The types and quantities of MPs found in different marine species are indicated in Table 3.

Many types of organisms are represented in these studies, including zooplankton, polychaete, bivalve, crustacean, fish, seabirds and mammals. Fish and seabirds have been the most widely studied organisms in marine ecosystems. Concerning the composition of the MPs found in the biota, information remains with only 3 studies having been published on PE, PP or polyamide (Moore et al., 2001; Lusher et al., 2013; Besselinga et al., 2015). The MP forms were widely indicated with filaments/fibers having been found, mostly in less evolved organisms: from zooplankton to Thaliacea, with and fragments in more evolved organisms from fish to mammals. As for water and sand/sediment samples, the units of the measured variables have differed between studies. The units were particles per gram or per individual or a proportion (%) of individuals with MPs recovered in part of the digestive tract, which does not allow evaluation of quantitative contamination in organisms. MPs were detected in many organisms as mentioned in Table 3 and detailed in the following paragraph.

Regarding the species of zooplankton in the Northeast Pacific Ocean, Desforges et al. (2015) reported MP quantity in Neocalanus cristatus and Euphausia pacifica. They showed that MPs were probably filtered by these species, the ingestion rate being dependent on their feeding mode, with MPs being found particularly in non-selective feeding species.

MPs found in the fecal casts of two polychaete species (*Clymella torquata* and *Alitta virens*) were equivalent to those found in sediment samples suggesting that ingestion of MPs equals egestion by the polychaete (Mathalon and Hill, 2014).

As a well-known bioindicator for environmental pollution in marine species, the mussel *Mytilus edulis* is the most widely studied group among marine organisms (4 studies). Low quantities of MPs were found in bivalves from German and French/Belgian/Dutch farms (respectively 0.36 ± 0.07 and 0.2 ± 0.3 particles/g of soft tissues; Van Cauwenberghe and Janssen, 2014; Van Cauwenberghe et al., 2015). Mathalon and Hill (2014) found greater quantities (up to 500 times higher) of MPs in the same species from Canada than those previously reported in bivalves from European areas by Van Cauwenberghe and Janssen (2014), Van Cauwenberghe et al. (2015). These discrepant results could be due to the differences in contamination levels between the sites studied. Moreover, as discussed above, the results also depend on the method employed for MP extraction (digestion and filtration) and analysis (microscopy and spectroscopy). The gut depuration step that was performed by Van Cauwenberghe and Janssen (2014), but which is not mentioned in the work of Mathalon and Hill (2014), could lead to differing results, since a decrease of up to 33% in the quantity of MPs in mussels appeared after 3 days of gut depuration (Van Cauwenberghe and Janssen, 2014). Comparison between commercial and wild types of mussels (*M. edulis*, *Mytilus galloprovincialis*) showed that the total number of MPs was not significantly different between the commercial mussels and wild specimens, with values varying from 2.6 to 5.1 fibers (0.2–1.5 mm in length) per 10 g of mussel tissues (De Witte et al., 2014). Regarding another bivalve, 0.47 ± 0.16 particles/g were found in the soft tissue (wet weight) of oysters (*Crassostrea gigas*) (Van Cauwenberghe and Janssen, 2014) with a decrease of approximately 25% after 3 days of depuration in clean seawater.

In a crustacean (*Nephrops norvegicus*), 83% of the animals sampled in the Clyde Sea contained plastic filaments in their stomach (Murray and Cowie, 2011). In this study, tightly tangled balls of plastics were predominantly found in up to 62% of the
<table><caption>**Table 3**
Type of organisms, species, type or form and quantity of MPs found in biota with references of the corresponding studies.</caption><thead><tr><th>Organisms</th><th>Species</th><th>Microplastics type or form</th><th>Quantity</th><th>References</th></tr></thead><tbody><tr><td rowspan="2">Zooplankton</td><td>Neocalanus cristatus</td><td>Fibers (50%)</td><td>0.026 particles/individual</td><td>Desforges et al. (2015)</td></tr><tr><td>Euphausia pacifica</td><td>Fibers (68%)</td><td>0.058 particles/individual</td><td></td></tr><tr><td rowspan="2">Polychaete</td><td>Clymenella torquata</td><td>Fibers</td><td>2-8/gram (equal in sediment)</td><td>Mathalon and Hill (2014)</td></tr><tr><td>Alitta virens</td><td>Fibers</td><td>2-8/gram (equal in sediment)</td><td></td></tr><tr><td rowspan="2">Bivalvia</td><td>Mytilus eduis</td><td>Fibers</td><td>106-178/mussel<br>0.36 ± 0.07 particles/g</td><td>Mathalon and Hill (2014)<br>Van Cauwenberghe and Janssen (2014)</td></tr><tr><td>Mytilus galloprovincialis<br>Crassostrea gigas</td><td>Fibers</td><td>0.2 ± 0.3 particles/g<br>0.26–0.51/gram</td><td>Van Cauwenberghe et al. (2015)<br>De Witte et al. (2014)</td></tr><tr><td>Crustacean</td><td>Nephrops norvegicus</td><td>Filaments</td><td>0.47 ± 0.16 particles/g</td><td>Van Cauwenberghe and Janssen (2014)</td></tr><tr><td rowspan="2">Thaliacea</td><td>Crangon crangon</td><td>Synthetic fibers</td><td>83% had MPs in stomach</td><td>Murray and Cowie (2011)</td></tr><tr><td>Thetys vagina</td><td>Fragment, PP/monofilament</td><td>0.68 ± 0.55 MP/g</td><td>Devriese et al. (2015)</td></tr><tr><td>Fish</td><td>Astronesthes indopacifica<br>Cololabis saira<br>Hygophum reinhardtii<br>Loweina interrupta<br>Myctophum<br>aurolanternatum<br>Symbolophorus<br>californiensis<br>Merlangius merlangus</td><td>Fragments</td><td>MPs predominant</td><td>Moore et al. (2001)<br>Boerger et al. (2010)</td></tr><tr><td rowspan="10">Seabird</td><td>Micromesistius poutassou</td><td>Polyamide (35.6%), polyester (5.1%), PS (0.9%), LDPE (0.3%)</td><td>32% fish had MPs<br>51.9% fish had MPs<br>28.6% fish had MPs<br>40% fish had MPs<br>47.6% fish had MPs<br>51.5% fish had MPs<br>38% fish had MPs<br>32.3% fish had MPs<br>26% fish had MPs<br>23.5% fish had MPs</td><td>Lusher et al. (2013)</td></tr><tr><td>Trachurus trachurus</td><td></td><td></td><td></td></tr><tr><td>Trisopterus minutus</td><td></td><td></td><td></td></tr><tr><td>Zeus faber</td><td></td><td></td><td></td></tr><tr><td>Aspitrigla cuculus</td><td></td><td></td><td></td></tr><tr><td>Callionymus lyra</td><td></td><td></td><td></td></tr><tr><td>Cepola macrophthalma</td><td></td><td></td><td></td></tr><tr><td>Buglossisium luteum</td><td></td><td></td><td></td></tr><tr><td>Microchirus variegates</td><td></td><td></td><td></td></tr><tr><td>Puffinus tenuirostris</td><td>Fragments</td><td>Observation</td><td>Tanaka et al. (2013)</td></tr><tr><td>Fulmarus glacialis</td><td>Fragments</td><td>Observation</td><td>Bond et al. (2014)</td></tr><tr><td>Puffinus gravis</td><td></td><td>Observation</td><td></td></tr><tr><td>Puffinus griseus</td><td></td><td>Observation</td><td></td></tr><tr><td>Calonectris diomedea</td><td>Fragments</td><td>Observations</td><td>Codina-García et al. (2013)</td></tr><tr><td>Puffinus mauretanicus</td><td></td><td></td><td></td></tr><tr><td>Puffinus yelkouan</td><td></td><td></td><td></td></tr><tr><td>Morus bassanus</td><td></td><td></td><td></td></tr><tr><td>Ichthyaetus audouinii</td><td></td><td></td><td></td></tr><tr><td>Ichthyaetus. melan officephalus</td><td></td><td></td><td></td></tr><tr><td>Larus michahellis</td><td></td><td></td><td></td></tr><tr><td>Rissa tridactyla</td><td></td><td></td><td></td></tr><tr><td>Catharacta skua</td><td></td><td></td><td></td></tr><tr><td>Mammalia</td><td>Phoca vitulina</td><td></td><td>11.2% seal had MPs in stomach<br>1% seal had MPs in intestine<br>16 MPs in gastrointestinal</td><td>Bravo Rebolledo et al. (2013)<br>Besselinga et al. (2015)</td></tr><tr><td></td><td>Megaptera novaeangliae</td><td>PE (56.3%)</td><td></td><td></td></tr></tbody></table>

animals, and there were no significant differences in plastic load between males and females. Devriese et al. (2015) found fibers in 63% of the shrimp studied (Crangon crangon), with an average value of 0.68 ± 0.55 MP/g wet weight.

The only species studied in Thaliacea was the salp (Thetys vagina), which presented plastic fragments and PP/monofilaments line firmly embedded in their tissues (Moore et al., 2001).

The same methodology as Moore et al. (2001) was used by Boerger et al. (2010) to determine the quantity of MPs in the guts of various common planktivorous fish from the North Pacific Central Gyre. About 35% of the sampled fish contained MPs in their guts (N = 670, 6 different species). The average number of MPs was 2.1 ± 5.78 pieces/fish. Furthermore, these authors showed that there was a positive correlation between size class of fish and quantity of MPs in the guts. The presence of MPs in the gastrointestinal tracts of pelagic and demersal fish from the English Channel was also investigated by Lusher et al. (2013). The authors studied 10 different species (5 pelagic and 5 demersal fish;

N = 504). A proportion of 36.5% (184/504) of fish contained MPs in their gastrointestinal tracts. The average number of pieces of plastic per fish was 1.90 ± 0.10 (351 pieces in 184 fish). Among the total of 351 plastic pieces 92.4% were MPs. There were no differences between pelagic and demersal fish.

MPs were also found in many species of seabirds (Bond et al., 2014; Codina-Garcia et al., 2013; Tanaka et al. 2013).

Bravo Rebolledo et al. (2013) showed that younger animals phoca vitulina accumulated greater quantities of MPs compared to older ones, and MPs were found mainly in the stomach (11.2%; N = 107) and the intestines (1%; N = 100) but not in the feces (N = 125).

Very recently, MPs were observed in the intestines of baleen whales (Megaptera novaeangliae), a macro-filter-feeder (Besselinga et al., 2015). A total of 45 particles were extracted from their intestines; 16 particles were identified as MPs of many different types: PE, nylon, PP etc. Among them, the most abundant type (56.25%) was PE, which is the most widely produced polymer
worldwide, and also the most frequently used in fishing gear.

## 2.4. Characterization and identification of MPs in field samples

The nature and quantities or concentrations of MPs were not systematically determined in field samples (Tables 1–3) due to the fact that improved identification of MPs from environmental samples requires several steps and demands mastery of several techniques. This is particularly the case for MPs from marine organisms. Differences in MP distribution due to changes of analytical procedure from extraction to visual sorting and identification by spectroscopy have previously been discussed in the literature (Rocha-Santos and Duarte, 2015). However, as shown in Tables 1 and 2, once when identification was carried out, PE fragments were clearly shown to be the most frequently present polymer in both waters and sand/sediment samples. Seawater samples usually contained greater proportions of low density polymers such as PE and PP as has been reported in many studies (Law et al., 2010; Moore et al., 2001; Reisser et al., 2013; Van Cauwenberghe et al., 2015). Surprisingly enough, in sediment samples some studies have also reported high proportions of PE and PP in the whole MPs detected whereas their density in water is lower (Carson et al., 2011; Gregory, 1977, 1983; Khordagui and Abu-Hilal, 1994; Ng and Obbard, 2006; Turner and Holmes, 2011; Van Cauwenberghe et al., 2015; Vianello et al., 2013). This observation tends to support the hypothesis of increased MP density by colonization of a wide range of organisms (micro-organisms, algae, invertebrates etc.) and sorption of organic pollutants. In other studies, MPs with high density (higher than seawater) were present in a more sizable proportion in sand/sediment samples compared to the water column (Browne et al., 2010; Claessens et al., 2011; Lee et al., 2013). Browne et al. (2010) showed that PVC represented 26%, polyester 35%, polyamide 18% of the MPs found in sediment samples. Claessens et al. (2011) also showed that PVC, polyester and polyamide comprised 80% of the total MPs in sediment samples from Belgian coastal areas. Lee et al. (2013) found that more than 96% of the plastic debris on beaches of Korea were Styrofoam (expanded PS) originating in oyster aquaculture facilities near the sampling sites and being used to sustain buoyancy with oyster growth, thereby showing that the nature and concentration of MP may also be dependent on the presence of industries in the vicinity of the sampling site. For deep-sea sediment samples, an average proportion of 53.4% for plastic fibers (polyester) was reported by Woodall et al. (2014) in the Northeast Atlantic Ocean, Southwest Indian Ocean and the Mediterranean Sea.

Considering their forms, plastic fragments were predominant in most of the studies in water and sand/sediment samples (Doyle et al., 2011; Eriksen et al., 2013; Ivar Do Sul et al., 2009, 2013; Law et al., 2010; McDermid and McMullen, 2004; Moore et al., 2002). Nevertheless, in some cases, other forms such as fibers and/or filaments were detected depending on the sampling method, the method of analysis, and the specific objective of the studies (Collignon et al., 2012; Desforges et al., 2014; Gregory, 1983; Lusher et al., 2014; Thompson et al., 2004; Woodall et al., 2014 and Zhao et al., 2014). Concerning MPs detected in organisms, the form was shown to depend upon animal with several observations of filaments/fibers in species from zooplankton to Thaliacea and of fragments in species from fish to mammals (Table 3).

Concerning the size of MP, Morét-Ferguson et al. (2010) reported that 60% of the plastic debris sampled within the North Atlantic subtropical gyre are 2–6 mm long. In addition, the change of size and abundance of MPs between the 1960–1970s and the 1980–1990s was studied by Thompson et al. (2004) who observed decreased size parallel to increased abundance, suggesting a phenomenon of fragmentation. As for Lee et al. (2013), they showed

that MP abundance depended on season with a threefold increase after as compared with before the rainy season, an increase suggesting a more intense fragmentation of MPs by the high physical energy appearing during typhoons.

Interspecies differences (e.g. specific biological barriers), means of exposure and MP abundance, type, size and form present in the field (water and sediment) evidently lead to different bioaccumulations. For example, some species filter what they ingest; others choose MPs according to their size, color etc. due to their resemblance to their usual food (Boerger et al., 2010). Moreover, depending on the object studied (organisms, water or sediment), the MPs under consideration are not the same. In organism studies, focus has typically been placed on small MPs like filaments, largely because they were the ones most widely found. In environmental studies, focus on bigger MPs has been easier with the use of manta nets or water filtration techniques with large pore sizes (>330 µm) to prevent clogging. Lots of MPs present in the environment have not been observed in organisms because they did not pass through the biological barriers or due to the egestion of the organisms; this phenomenon has been reported in many studies (Van Cauwenberghe and Janssen, 2014; Mathalon and Hill, 2014). Consequently, the link between what is found in organisms and what is found in the environment is complex.

Many different analytical techniques are currently used to assess the composition and concentration of MPs; they include Fourier-transform infrared spectroscopy (FT-IR, Thompson et al., 2004; Vianello et al., 2013; Nor and Obbard, 2014), Raman spectroscopy (Van Cauwenberghe and Janssen, 2014, 2015) and pyrolysis-gas chromatography-mass spectrometry (Pyr-GC-MS, Fries et al., 2013; Dekiff et al., 2014). Their principles are different hence the results may differ as well. As regards MPs in marine wildlife, the use of chemical digestion for extraction from organisms has seemed to influence results. For example, nitric acid extraction efficiency depends on MP sizes and natures. Efficiency is 30% higher for particles of 30 µm compared to 10 µm particles whereas the nylon rope fibers are not extracted (Claessens et al., 2013). There is consequently an urgent need for standardization of the methods for collection, extraction and analysis of MP contamination in the field and animals.

## 3. Exposure experiments in the laboratory

Over recent years, experimentations designed to assess the ingestion and biological effects of MPs have been carried out on many species, ranging from zooplankton to birds (Table 4). The highest number of studies have involved polychaetes (only one species: *Arenicola marina*) and crustaceans whereas zooplankton has been the most widely investigated in terms of number of species. Concerning the type of MPs used in exposure experiments, the majority of the studies have been performed with PE or PS and, exceptionally, with PVC and PP. Most of the time, the shapes of MPs were spherical since the sphere, often known as a microbead, is the most commonly commercialized shape. The size of the MPs has depended on the study under consideration as well as the concentrations/quantities of MPs introduced in exposure media. The concentrations/quantities used for exposures are very variable and tend to be expressed in different units, leading to difficulties for further comparisons. For example, the units employed have been particle number/medium volume, particle mass/medium volume, particle number/organism number, particle number/sediment mass, etc. Exposure time has also greatly varied, ranging from 4 h to 2 months. Consequently, in addition to the above-mentioned difficulty arising from the use of different unit expressions of MP quantities, variations in MP sizes, quantities and exposure times have led to highly problematic comparisons between the results of
<table><caption>**Table 4**
Type of organisms, species, type, form and concentrations/abundance of MPs as well as time conditions of laboratory exposures with references of the corresponding studies.</caption><thead><tr><th>Organisms</th><th>Species</th><th>Microplastics(MPs) Type* or form</th><th>MP Concentrations/abundance</th><th>Exposure time</th><th>References</th></tr></thead><tbody><tr><td>Zooplankton</td><td>Marenzelleria spp<br/>Acartia spp<br/>Limnocalanus macrurus<br/>Eurytemora affinis<br/>Synchaeta spp<br/>Tintinnopsis lobiancoi<br/>Neomysis integer<br/>Mysis relicta<br/>Mysis mixta<br/>Bosmina coregoni<br/>nordmannii<br/>Evadne nordmannii<br/>Oxyrrhis marina<br/>Doliolidae<br/>Euphausiidae Obelia sp<br/>Siphonophorae<br/>Bivalvia<br/>Brachyura<br/>Caridea<br/>Paguridae<br/>Porcellanidae<br/>Mytilus edulis</td><td>PE sphere 10 µm<br/>PS 7.3–30.6 µm</td><td>1000–10000 particles/mL<br/>635–3000 beads/mL</td><td>12 h<br/>24 h</td><td>Setälä et al. (2014)<br/>Cole et al. (2013)</td></tr><tr><td>Bivalvia</td><td>Mytilus galloprovincialis<br/>Arenicola marina</td><td>HDPE 0–80 µm no additive<br/>PS 3–9.6 µm<br/>PS 0.5 µm<br/>PS microsphere (10–90 µm)</td><td>2.5 g/L<br/>42 particles/mL<br/>50,000,000 particles/mussel<br/>110 particles/mL</td><td>96 h<br/>To 48 days<br/>1 h<br/>14 days</td><td>Von Moos et al. (2012)<br/>Browne et al. (2008)<br/>Farrell and Nelson (2013)<br/>Van Cauwenberghe et al. (2015)<br/>Avio et al. (2015)<br/>Besseling et al. (2013)</td></tr><tr><td>Polychaeta</td><td>Holothuria floridana<br/>Holothuria grisea<br/>Thyonella gemmata<br/>Cucumaria frondosa<br/>Embryo of Lytechinus variegatus<br/>Centropages typicus<br/>Calanus helgolandicus<br/>Acartia clausi<br/>Temora longicornis<br/>Allorcheses Compressa<br/>Orchestia gammarellus<br/>Semibalanus balanoides<br/>Nephrops norvegicus<br/>Calinus maenas</td><td>PE and PS (0.1–1 mm)<br/>PS sphere (400–1300 µm) pre-equilibrated with PCBs<br/>PVC pre-equilibrated POPs (NP*, Phe*, PBDE*, Triclosan)<br/>PE pre-equilibrated with Phe<br/>MP 20–2000 µm<br/>PS microsphere (10–90 µm)<br/>PVC 130 µm<br/>MP collected on beach<br/>PVC (fragment and pellet), Nylon <5 mm<br/>PE<br/>PS 0.4–30.6 µm<br/>PE 11–700 µm<br/>MP 20–2000 µm<br/>PP rope<br/>PS 0.5 µm</td><td>MP = 200–260 ng/g<br/>0, 0.074, 0.74 and 7.4% dry weight<br/>PVC 5% in sand<br/>Lower ratio in sediment<br/>1.5 mg MP/individual<br/>110 particles/g sediment<br/>0–5% in sediment<br/>16.7 g/L<br/>solution MP/solution exposure: 20%<br/>635–3000 beads/mL</td><td>7 days<br/>28 days<br/>10 days<br/>Few days<br/>14 days<br/>48 h<br/>4 h<br/>24 h<br/>24 h</td><td>Teuten et al. (2007)<br/>Thompson et al. (2004)<br/>Van Cauwenberghe et al. (2015)<br/>Wright et al. (2013)<br/>Graham and Thompson (2009)<br/>Nobre et al. (2015)<br/>Cole et al. (2013)</td></tr><tr><td>Crustacean</td><td>Calanus helgolandicus<br/>Pomatoschistus microps<br/>Oryzias latipes</td><td>PS 7.3–30.6 µm<br/>PS 20 µm<br/>PE 420–500 µm<br/>PE 1–5 µm<br/>LDPE <0.5 mm</td><td>635–1,000,000 beads/mL<br/>0.1 g MP/individual<br/>1 g MP/individual<br/>1 g MP/1 lit/10 individual<br/>Eat fish pre-contaminated with MP<br/>Eat mussels pre-contaminated with MP<br/>75 MP/mL<br/>7 types of prey<br/>18.4–184 µg/L<br/>10% of their prey</td><td>72 h<br/>Few days<br/>24 h<br/>To 21 days<br/>24 h<br/>24 h<br/>To 96 h<br/>2 months<br/>To 42 days</td><td>Chua et al. (2014)<br/>Thompson et al. (2004)<br/>Murray and Cowie (2011)<br/>Farrell and Nelson (2013)<br/>Cole et al. (2015)<br/>De Sa et al. (2015)<br/>Oliveira et al. (2013)<br/>Rochman et al. (2013b, 2014)<br/>Teuten et al. (2009)</td></tr><tr><td>Seabird</td><td>Calonectris leucomelas</td><td>PE coastal</td><td>1.3–1.4 g/bird</td><td>To 42 days</td><td></td></tr></tbody></table>

different studies. Most works have focused on the ingestion and biological effects of MPs after laboratory exposure of animals. Nevertheless, in some studies trophic transfer has been studied with regard to exposure of organisms to preys previously contaminated with MPs (Farrell and Nelson, 2013; Murray and Cowie, 2011).

## 3.1. Ingestion of MPs by organisms

All of the results of marine organism exposures to MPs, from zooplankton to birds, highlight some common trends. The ingestion of MPs has depended on several different factors.

The first factor is the influence of the species studied. For example, concerning zooplankton, several species originating from the Baltic Sea were collected and their capacity of MP ingestion was

<footnote>(*): HDPE: High Density Polyethylene; LDPE: Low Density Polyethylene; NP: Nonylphenol; PBDE: Polybrominated diphenyl ether; PE: Polyethylene; Phe: Phenanthrene; PS: Polystyrene; PVC: Polyvinyl chloride.</footnote>
tested by Setälä et al. (2014). These authors exposed 11 species to 10 µm fluorescent PS microspheres at 3 concentrations (1000, 2000, 10,000 particles/mL) during 12 h. While 9 out of the 11 species ingested MPs, appreciable differences in the percentages of MP ingestion were observed. The highest percentage with ingested spheres was found in the pelagic polychaete larvae, *Marenzelleria* spp. In another study (Cole et al., 2013), 15 zooplankton taxa, which are representative of abundant mesozooplankton from Northeast Atlantic coastal systems, were collected in Plymouth, UK. PS spheres were chosen in the size range of 7.3–30.6 µm for the purposes of laboratory exposure during 24 h. Their concentrations ranged from 635 to 3000 beads/mL, inversely correlated to the size of the beads. As a result, 13 out of 15 exposed species showed MP ingestion at different proportions depending on taxa. Graham and Thompson (2009) studied four species of deposit- and suspension-feeding sea cucumbers (*Holothuria floridana*, *Holothuria grisea*, *Thyonella gemmate* and *Cucumaria frondosa*) exposed to PVC pellets (4.0 mm) and showed that the large size of the pellets likely prevented the two smallest species (*Thyonella gemmata* and *H. floridana*) from ingestion, thereby suggesting that feeding capacity could depend on species size.

Intra-species differences could also be a factor influencing the ingestion of MPs. Farrell and Nelson (2013) exposed *M. edulis* to fluorescent PS microspheres of 0.5 µm for 1 h (411 million particles for 8 mussels × 3 replicates). In 13 out of 24 mussel organisms, the microspheres were sufficiently numerous to be visible to the naked eye when cut open. Moreover, for seabirds (*Calonectris leucomelas*, *Procellaria aequinoctialis*), ingestion of MPs was demonstrated by Ryan and Jackson (1987) and Teuten et al. (2009). Age appeared to be a major factor in the contamination by MPs since young and immature birds were more highly contaminated than older ones. Nevertheless, the author did not find any hypothesis supporting this observation (van Franeker and Law, 2015).

In addition to inter- and intra-species differences, another factor influencing ingestion could be MP size. Cole et al. (2013) showed that at both life-stages, adult and juvenile, the Holoplankton *Calanus Helgolandicus* ingested MPs of 20.6 µm while the Holoplankton *Acartia clausi* adults ingested MPs 30.6 µm MPS but not 20.6 µm MPs. Browne et al. (2008) showed that the accumulation of smaller particles (3 µm) by *M. edulis* was superior to 60% compared to larger particles (9.6 µm), in accordance with results from Van Cauwenberghe et al. (2015) who carried out mussel exposures to PS of three different sizes (10, 30, 90 µm). As a result, the accumulation in the tissues of lugworms of 10 µm particles was greater than 30 µm particles whereas no accumulation was shown for 90 µm particles.

Type of MP is also likely to affect MP ingestion and could be an influencing factor. The four species of deposit- and suspension- feeding sea cucumbers (*H. floridana*, *H. grisea*, *T. gemmate* and *C. frondosa*), studied by Graham and Thompson (2009) were exposed during 4 h of daylight to three types of plastic: PVC fragments (0.25–15 mm long, most of them smaller than 5 mm), nylon fragments (0.25–15 mm long, 0.27 mm in diameter), and PVC pellets (4.0 mm in diameter) all of them collected along the east coast of the USA from Florida to Maine. All of the species ingested MPs in variable proportions. The researchers observed significantly more sizable ingestion of nylon and PVC fragments than pellets (2–138 fold), suggesting selective ingestion by organisms.

Only one study has examined the effect of MP color after laboratory exposure and it seemed, in this case, that it was a factor relevantly influencing MP ingestion. Juvenile fish *Pomatoschistus microps* were exposed during 24 h to PE microspheres (420–500 µm, 3 colors: white, black and red) (De Sa et al., 2015). The authors observed referential ingestion of white as opposed to blue or red MPs. They then went on to suggest that MPs could be

confused with food.

Most of the time, when ingested MPs have been localized within an organism, the digestive tract has been mentioned, according to a study of Von Moos et al. (2012) that highlighted the presence of HDPE (0–80 µm) in the digestive system of *M. edulis* after 3 h of exposure. Concerning other animals, two filter-feeder species, the barnacle *Semibalanus balanoides* and the amphipod *Orchestia gammarellus* from intertidal habitats close to Plymouth (UK) were exposed to MPs (20–2000 µm) at high concentrations (1 g MPs/L of seawater/10 individuals and 1 g MPs/individual respectively) for a few days (Thompson et al., 2004). Microscopic observation revealed the presence of MPs in the intestinal tracts of both species.

As mentioned above, several factors were identified as having or not having an impact on MP ingestion by organisms, e.g. inter- and intra-species differences, size, type and color of MPs. Combinations of these factors probably occur, which renders it difficult to fully comprehend MP ingestion. Nevertheless, after ingestion by organisms, MPs can potentially be transported in the whole body, leading to the biological effects investigated in exposure studies (Table 4) and discussed in the following section.

## 3.2. Biological effects

The biological effects of MPs have been explored in different invertebrate and vertebrate species after laboratory exposures (Von Moos et al., 2012; Oliveira et al., 2013; Cole et al., 2013, 2015). Among the different biological effects, mortality rate, energy budget, loss of weight, feeding activity, embryonic development, predation, biomarker responses and alteration of gene expression have been the most investigated.

The hypothesis that MPs are taken up into cells and can cause significant effects on tissue and at the cellular level was corroborated by Von Moos et al. (2012) in mussels (*M. edulis*). Through the exposure of mussels to high-density additive-free polyethylene (HDPE), with non-uniformly shaped grains ranging from 0 to 80 µm and at a nominal concentration of 2.5 g/L, MPs led to uptake in gills followed by transport towards the stomach and the digestive gland, where accumulation occurred in the lysosomal system after 3 h (Von Moos et al., 2012). In this study, pronounced destabilization of lysosomal membrane and strong inflammatory responses were shown to occur when exposure time was increased. Browne et al. (2008) showed in mussels (*M. edulis*) that while ingestion and translocation of MPs did not change the phagocytic activity, it increased immune response. Concerning fish, specimens of *Pomatoschistus microps* were exposed to PE microspheres (1–5 µm) at concentrations ranging from 18.4 to 184 µg/L (Oliveira et al., 2013). After 96 h of exposure, a reduction of acetylcholinesterase (AChE) activity had been shown to occur. In contrast, no significant effect of PE was found for glutathione S-transferase activity and lipid peroxidation. Rochman et al. (2014) mixed low-density PE and preys of another fish species (*Oryzias latipes*) at a high proportion (up to 10% of the prey species) during 2 months. Several negative effects appeared: down regulation of choriogenin, vitellogenin and estrogen receptor (ERα) mRNA gene expression and abnormal germ cell proliferation. Severe glycogen depletion and fatty vacuolation were also observed. In the long term, a potential increase of mortality due to the effects observed at molecular level is still under debate. For example, Rochman et al. (2013, 2014) showed a mortality rate reaching 6% and probably associated with the effects observed at the molecular level. In contrast, Browne et al. (2008) showed that in mussel (*M. edulis*), exposure to PS microspheres did not affect their viability.

The survival of organisms is also ensured by other processes and can be impaired by a modified energy budget. While Van Cauwenberghe et al. (2015) did not measure any changes in the
energy budget of mussels (*M. edulis*) ingesting PS microspheres, an increase of energy consumption in the digestive gland (25%) was nonetheless observed. At high concentrations of exposure (up to 5% by weight, in sediment), Wright et al. (2013) showed a depletion of energy reserves (up to 50%) in lugworms (*Arenicola marina*), after 10 days of exposure, whereas despite longer exposure time (up to 14 days), Van Cauwenberghe et al. (2015) showed no depletion of energy reserves for this species at low concentrations. An energy budget depends not only on energy consumption but also, at a higher (eco)toxicologic level, on feeding activity. In one example, the impact of MPs on copepod (*Centropages typicus*) feeding activity was investigated by Cole et al. (2013, 2015). A significant decrease of algal feeding was shown under different conditions of MP exposure (>4000 beads of PS 7.3 µm/24 h and 75 beads of PS 20 µm/ 24 h, Cole et al., 2013, 2015 respectively). For lugworms (*Arenicola marina*) exposed to MPs, a reduced feeding activity was likewise shown in two different studies (Besseling et al., 2013; Wright et al., 2013). Concerning some vertebrate species, predation can also be studied. De Sa et al. (2015) showed that predation of a fish species (*Pomatoschistus microps*) and its efficiency were reduced by 65% and 50% respectively in the presence of PE microspheres with a diameter of 420–500 µm and different colors. A loss of weight in *A. marina* was indeed observed when MP concentration increased in exposure media (Besseling et al., 2013). Consequently, we could assume that the reduction of predation as well as the abundance of MPs in the media modify energy consumption, which is directly linked to the energy budget and loss of organism weight, which could lead to reduce predation, thereby completing a cycle.

At the ecological level, population survival was demonstrated to be dependent upon organism survival, reproductive efficiency, embryonic development, etc. These endpoints were investigated in the following three studies. Although there was no significant effect of MP exposure on production rates and egg size of the copepod (*Centropages typicus*), following exposure to MPs the hatching of eggs seemed depleted (Cole et al., 2013, 2015). The toxicity of PE on the embryonic development of an Echinodermata (*Lytechinus variegatus*) was also demonstrated by Nobre et al. (2015). After 24 h of exposure, PE pellets had negative effects on embryonic development, which was assessed in terms of the presence of abnormal embryos.

### 3.3. Trophic transfer

In addition to MP exposure from the media, trophic transfer has been studied in works involving different levels of the food web. The aim of these studies was to highlight MP intakes occurring through diet in addition to ingestion from the media (water and sediments). Trophic transfer is a process leading to a potential human health risk because humans are at the top of the trophic chains and consequently have got to be taken into account in research studies.

Potential MP transfer via planktonic organisms to a higher trophic level was indicated by MP observations in *Mysida* intestine after 3 h of incubation with mesozooplankton previously exposed to PE microspheres (Setälä et al., 2014).

Farrell and Nelson (2013) observed MP trophic transfer from mussels to crabs. *M. edulis* were exposed to 0.5 µm fluorescent PS microspheres (411 million particles for 8 mussels) during 1 h. After that, their tissues were cut and placed in a bucket with crabs (*Carcinus maenas*). Crabs were sampled from 1 h to 21 days. Microspheres were found in the stomach, the hepatopancreas, the ovary, the gills and the haemolymph of the exposed crabs, in decreasing numbers over the trial period.

Murray and Cowie (2011) worked on the crustacean *N. norvegicus* originating from the Clyde Sea. After cleaning the guts of

animals for 2 weeks, they were fed with their prey (fish) and pre-exposed with blue propylene rope (5 mm) for 24 h. The authors showed that 100% of the animals fed with plastic seeded fish presented plastics in their stomachs whereas 70% of the control animals contained plastics which they had accumulated prior to being captured and had not digested during the 2-week starvation period prior to the experiment. These results could be worrisome since this species has a strong socio-economic interest and is consumed by humans.

Another edible species was studied by Desforges et al. (2015), who estimated that consumption of the MPs contained in zooplankton led to the ingestion of 2–7 MP particles/day by members of the juvenile salmon species (*Oncorhynchus* spp.) from coastal British Columbia, and ≤91 MP particles/day in returning adults.

Finally, Van Cauwenberghe and Janssen (2014) estimated that annual dietary exposure for European shellfish consumers can reach 11,000 MPs per year. These results pose a challenge about consequences on human health.

## 4. Discussion

### 4.1. Are experimental laboratory exposures consistent with the exposome?

Many studies have focused on the MPs found on the field. Given the constraints associated with the sampling protocols, sample treatments and analytical methods, these field studies are hardly comparable. In this part, the results on MPs from field studies will be compared when possible with the conditions used for laboratory exposures, in order to determine the degree of consistency between laboratory exposure and field exposome. Table 5 is a comparative table to summarize the differences between the MPs found in the field and those used in laboratory exposures.

Concerning MPs in water, field concentrations expressed in particles/km² cannot be used for comparison with laboratory exposure, which is usually expressed as particles/medium volume, particle mass/medium volume, particle number/organism number, and particles/sediment mass. Only the expression of field concentrations in particles/m³ can be compared to exposures expressed in particles/medium volume. The range found in the field was 0.004–9200 particles/m³ (Table 1). In laboratory exposures, the contamination range expressed in particles/mL was 42 to 10,000 corresponding to 42 million to 10 billion particles/m³. Comparing these values, it is obvious that the concentrations were not of the same order of magnitude, the lowest concentration of exposure being about 4500 times greater than the highest field concentration.

In the literature, few works have studied the accumulation and effect of MPs on organisms after sediment exposures. The reported MP exposure concentrations were 5% in mass of sand (Browne et al., 2013), up to 5% in mass of sediment (Wright et al., 2013) and 110 particles/g of sediment (Van Cauwenberghe et al., 2015). In field studies, a majority of concentrations were expressed as particle number per m³, m² or L, rendering impossible any comparison with the concentrations usually employed in laboratory exposures. Some field MP concentrations were expressed in particle number/ kg of sediment with values ranging from 0.3 to 8000 corresponding to 0.0003 to 8 particles/g (Table 2). This value is more than 10 times below the concentration employed by Van Cauwenberghe et al. (2015). Only one study on natural sediments from the Indian Ocean (Reddy et al., 2006) has expressed the concentration as 81.43 mg/kg, corresponding to 0.0081%, which was about 600 times lower than the concentrations used by Browne et al. (2013) and Wright et al. (2013) in laboratory exposures.
<caption>**Table 5**</caption>

Overall comparison of the MPs found in waters and sediments with the different MPs used for laboratory exposures. The number of references citing this type or size or nature of MP as predominant is indicated in brackets. In gray: PVC; in black: PE; in blue: PP; in yellow: PS.* identification of MP nature was not performed in all references.

<table><thead><tr><th></th><th>Compartment</th><th>Type</th><th>Size</th><th>Nature</th><th>Concentration</th></tr></thead><tbody><tr><td rowspan="2">In the field</td><td>Water (25)</td><td>Fragments predominant (16/25)</td><td>&gt; 330 µm</td><td>PE, PP, PS (6/25*)</td><td>0.004 to 9200 particles/m³</td></tr><tr><td>Sediment and sand (27)</td><td>Fragments predominant (15/27)</td><td></td><td>PE (10/27*)</td><td>0.3-2175 particles/kg</td></tr><tr><td rowspan="2">Laboratory exposures</td><td>Water (24)</td><td>Beads, spheres (21/24); MPs from environment (3/24)</td><td>500 nm to 2 mm; &lt; 100 µm (15/25)</td><td>PS, PE, PVC (21/25)</td><td>42.10<sup>6</sup> to 1.10<sup>12</sup> beads/m³</td></tr><tr><td>Sediment (1)</td><td>Spheres</td><td>130 µm</td><td>PVC</td><td>110 particles/kg</td></tr></tbody></table>

When mentioned (in 6 out of 25 studies), the type of MPs found in natural waters was PE and at times additional PP or PS. All of the waterborne exposure studies were conducted with PE or PS, with some exceptions for PVC and PP, which seems to be consistent with the field. As regards the MP types found in sediment field samples, despite their low density, PE was widely predominant when the MP type was mentioned, and was also sometimes associated with PP, PVC and PS. The polymers used for sediment exposures were PVC or PS, which do not predominate in the field. In laboratory exposures, even when the choice of polymers was consistent with those found in the field, medium contaminations were usually carried out with only one type of polymer, whereas the MPs present in the field are known to be a mix of many types of plastic. Nevertheless, the justification for use of PE or PS in laboratory exposures could be that they represent 36% of total production of plastics (Andrady and Neal, 2009). Moreover, PVC is likely to have been used in sediment exposures because it is denser than water and consequently flows and accumulates in sediment. As mentioned above, while PE with low density can be found in field sediments due to the colonization of organisms and adsorption of compounds, it remains difficult to mimic the phenomenon in laboratory experiments. The use of low density PE for sediment exposure would lead to transfer of MPs from the sediment to the water column, which could modify exposure modalities.

Whether the exposure studies were for waterborne or sediment-containing environments, in terms of forms most of the time the MPs employed for exposures, were commercialized microspheres with precise and regular sizes whereas in the field, the reported MPs presented high polydispersity in size with various irregular forms, such as fragments, fibers or filaments. These microspheres can mimic the primary MPs but not the secondary MPs that are predominant in the environment.

In order to take into account these numerous and important differences between exposure conditions and exposome, some authors (Graham and Thompson, 2009; Nobre et al., 2015) have used MPs extracted from environmental samples to carry out medium contaminations. In their findings, MPs from marine samples were more in line with exposome, i.e., MPs with a variety of types and of various sizes as well as irregular and different forms.

Moreover, as previously mentioned, MPs from marine environments could be colonized by organisms and constitute the sorption support for many organic compounds present in the environment, such as polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs), organochlorines (DDEs) ... To mimic organic compound sorption, Rochman et al. (2014) deployed PE pellets in marine areas during 3 months to obtain MPs more similar to those found in the environment. In addition, the levels of plastic additives, like phthalates or bisphenol A, in MPs found in the environment should be lower than commercialized plastic polymers since the time spent in marine water is likely to lead to a desorption of these compounds. In contrast, commercialized polymers are likely to contain higher levels of additives, except if they are polymers without additives, which are not currently commercially available in microsphere shapes.

## 4.2. Do the toxicity effects reported on a wide range of aquatic organisms, exposed to MPs in laboratories reflect environmental reality?

It remains difficult to conclude that experimental exposures are likely to mimic environmental conditions in terms of MP contamination. Since laboratory MP contaminations are not always consistent with environmental conditions the toxicity effects observed on organisms exposed to MPs in laboratory cannot reflect the reality encountered in the field.

All the laboratory experiments were performed with MP concentrations in greater orders of magnitude than those found in the field in order to depict the associated toxicity mechanisms. Consequently, the ingestion and associated effects observed in organisms corresponded to highly pronounced contaminant stress. Studies employing environmental MP levels are challenging since the available analytical tools do not yet permit identification of the biological effects occurring at low concentrations of exposure. However, in the next decades environmental MP loads are not likely to decrease since plastic, with its numerous advantages, is more and more massively used. In this context, the study of biological impacts on organisms after exposure to high concentrations of MP may be of considerable interest.
In addition to the problems associated with highly variable MP concentrations, it is difficult to differentiate and separately measure the mechanical and the chemical effects of MPs on organisms. It would be interesting to determine whether the toxicity effects are due to the presence of particles or else related to the polymer nature or to the MP size and form. For example, Avio et al. (2015) showed that the majority of the observed biological variations on exposed organisms (immunological, lysosomal, cholinesterasic and antioxidant effects) were not influenced by the type of polymer (PE vs PS). In order to determine the mechanical effects, the use of particles similar to MPs in laboratory exposures, except in their nature (glass particles for example), could be useful.

As it is now well-established that MPs play the role of vectors of organic compounds such as plastic additives or organic pollutants, to marine organisms (Teuten et al., 2007; Browne et al., 2013; Besseling et al., 2013), so as to create more environmentally realistic conditions, a few impact studies have used MPs associated with organic pollutants. The organic compounds include non-ylphenol, triclosan, pyrene, polybromodiphenylethers (PBDEs), PAHs, PCBs (Browne et al., 2013; Oliveira et al., 2013; Chua et al., 2014; Avio et al., 2015) which are known to cause toxic effects by themselves (Meeker et al., 2009; Oehlmann et al., 2009; Talsness et al., 2009; Vidal-Linan et al., 2015). Consequently, the presence of these compounds in MPs generated an additional effect, rendering it difficult to determine from where the toxicity arises. As a result, most studies have tried to differentiate the effects of the MPs from the effects of sorbed organic compounds through exposure experiments performed with MPs mixed or not mixed with organic compounds. In general, in these studies the different organisms exposed to virgin or contaminated MPs exhibited different effects, [lugworms (Arenicola marina), mussels (M. galloprovincialis), common goby (Pomatoschistus microps)], suggesting that impacts were induced by physical ingestion of the particles and the chemical toxicity of the adsorbed organic compounds.

In the case of Rochman et al. (2013, 2014), they used virgin pellets and pellets previously deployed during 3 months in marine urban areas for exposure to Japanese medaka (O. latipes). They reported that in a marine environment, PE MPs sorbed environmentally relevant concentrations of contaminants. They also found effects on gene expression (down-regulation of choriogenin) in males exposed to marine-plastic treatment, whereas gene expression (down-regulation of vitellogenin, choriogenin and the estrogen receptor) was altered in female fish exposed to both marine- and virgin-plastic treatment, suggesting that the ingestion of these plastic debris (and the associated chemical pollutants) at environmentally relevant concentrations may alter endocrine system function in fish.

In order to ensure more environmentally representative studies, the toxicity of two different types of PE, i.e., virgin or beach-stranded plastic pellets, on the embryonic development of an Echinodermata (L. variegatus) were studied by Nobre et al. (2015). The toxicity of stranded pellets was found to be lower than that of virgin pellets. As shown by Rochman et al. (2013, 2014), pellets from beach had greater concentrations of hydrophobic pollutants than did virgin pellets. Nevertheless, the authors suggested that plastic pellets acted as a vector of pollutants, especially for plastic additives found on virgin particles, and also suggested that the toxicity of leached chemicals from pellets could depend on the environmental compartment in which the pellets accumulated. This is the only study highlighting the effects of potential plastic additives, which is obviously highly important. While the choice of the nature of polymer could have no effect as observed by Oliveira et al. (2013), the use of raw polymers with or without additives has to be taken into account when biological effects are studied

through laboratory exposure.

## 5. Conclusion

There is an urgent need to establish standardized protocols for sampling, sample preparation, MP analysis and data expression so as to facilitate the comparison of field studies. It would be helpful to better define exposome in view of producing laboratory exposure conditions that would be more consistent with the natural environment. Guidelines should also be introduced to better mimic environmental conditions and translate into realistic biological effects. Many investigations have got to be continued: (i) study of organic compound sorption on MPs taking different abiotic factors into account, (ii) mechanical and chemical toxicity evaluation of exposures to MPs associated with organic compounds under realistic environmental conditions, (iii) evaluation of biological effects after exposure to MPs at different levels of the biological organization using several approaches such as “omic” methods (genomic, transcriptomic, metabolomic), biochemical and behavioral biomarkers. New investigations have also got to be carried on aimed at broadening the MP size ranges studied to “nanoplastics”, which are challenging because of a need for new and more efficient analysis methods. To conclude, MPs toxic for organisms are likely to be present in many marine environments, raising the risk of severe alterations to human health.

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<figure><img src="image_67.png" /></figure>
a wide range of biological activities including an antibacterial activity against Gram-positive bacteria [9–11], as well as dormant mycobacteria [12,13] and an antifungal activity [10]. They have also been widely studied for their potential use as biocontrol agents against plan fungal pathogens [14–16]. They have been reported to have an antiviral activity, particularly against infections caused by the virus of tobacco mosaic [17–19], and an antiparasitic activity against amoebae (Dictyostelium sp.) [20] and protozoa (Plasmodium falciparum) [21]. Peptaibols are supposed to play a role in the anti-phytopathogenic activity of Trichoderma sp. [15,22]. If they are long enough, they are able to form voltage-dependent ion channels in the lipid membranes [23,24]. Consequently, Trichoderma strains producing high amounts of long-chain peptaibols (18–20-residues) could represent potential new BCAs. Moreover, peptaibols could have not the same efficiency according to the targeted phytopathogen. It thus necessary to quantified separately each produced peptaibol subgroups in order to select the most efficient Trichoderma strain to inhibit the targeted pathogen.

Peptaibols could be classified according to their amino acids chain length and are produced in microheterogenous mixtures. The main subgroups are 11-, 14-, 18-, 19- and 20 amino acid residue peptaibols [25]. Until now, the LC–MS quantifications of peptaibols have been performed using alamethicin, as standard [26,27]. Currently, all peptaibol subgroups (11-res–20-res peptaibols) were quantified using only one standard: the commercial alamethicin mixture. Commercial alamethicin is a mixture of four 20-residue peptaibols, being the only current peptaibol commercialized mixture. Moreover, according to the study from Degenkolb, this commercial mixture contain a mycotoxin: the trichothecene [28] In the present study a LC/ESI-IT-MS methods was developed using different standards of peptaibols (11-, 14- and 20-residues) to quantify each peptaibol subgroup separately. It seems necessary to use standards with different chain lengths since they probably lead to different ionizations in the mass spectrometer (ESI-IT) and consequently to different sensitivities.

In the present study, after development and validation of the analytical methods of the quantification of the different peptaibol subgroups, 12 marine-derived Trichoderma strains of the marine fungal strain collection of the laboratory were screened in order to select the best potential BCAs.

## 2. Experimental

### 2.1. Chemical

Dichloromethane (CH₂Cl₂) and methanol (MeOH) used for peptaibol extractions and purifications were purchased from Carlo Erba (Val de Reuil, France) and distilled prior to use. The HPLC quality grade methanol for MS analyses was obtained from Biosolve (Valkenswaard, Netherlands).

### 2.2. Peptaibol standards

Alamethicin F50/5, hypomurocin A-III and a bergofungin D analogue were synthesized using microwave-assisted one shot synthesis according to the protocol of Ben Haj Salah and Inguimbert [29]. They are respectively named PEP-20 (20-residue peptaibols), PEP-11 (11-residue peptaibols) and PEP-14 (14-residue peptaibols) in this paper. Hypomurocin A-III (PEP-11) was observed in *Hyprocrea muroina* [30] and the similar sequence of bergofungin D (PEP-14) isolated from *Emericellopsis donezki* [31] contain pro-line residues in positions 9 and 12 instead of hydroxyproline. Their sequences are:

* PEP-11: AcUQVLUPULLUPLOl

* PEP-14: AcVUUVGLUUPQUPUFoI

PEP-20: AcUPUAUAQUVUGLUPVUUQQFol

* Q = Glutamine
* V = Valine
* U = Aib
* P = Proline
* L = Leucine

* A = Alanine
* Fol = Phenylalaninol
* Lol = Leucinol
* G = Glycine.

### 2.3. Fungal strains

Twelve marine-derived and one commercial strain of *Trichoderma* sp. were selected for this study. Their specificities are given in Table 1. Their qualitative peptaibol productions have been determined using mass spectrometry analyses in infusion mode as described by Carroux et al. [32].

### 2.4. Culture

Strains were inoculated onto Petri dishes (10-cm diameter) containing 20 mL of DCA medium (Dextrose 40 g/l, enzymatic digest of Casein 10 g/l, Agar 15 g/l, DIFCO, France) prepared with reconstituted seawater. Cultures were incubated in natural light at 27 °C.

### 2.5. Extraction and purification

After incubation, mycelia and conidia were scraped from the agar surface. The harvested biomass was steeped twice in CH₂Cl₂/MeOH 1:2 then 2:1 (v/v, 30 mL total volume) for 30 min at room temperature. The combined extracts were filtered, washed with distilled H₂O (20 mL) and evaporated to dryness to provide crude extracts.

### 2.6. LC–MS analysis

The samples were analyzed using a high performance liquid chromatography (HPLC) system consisting on a spectraSYSTEM P1000XR pump, a spectraSYSTEM AS3000 as autosampler (Thermo Fisher Scientific, San Jose, CA, USA) equipped with a Kromasil-C18 5 µm 250 mm × 2 mm column (Interchim, France) heated at 40 °C. The autosampler was maintained at room temperature and was programmed to inject 5 µL of sample in the chromatographic column. The mobile phase consisted on a MeOH/H₂O mixture (85:15, v/v) in isocratic mode at a constant flow rate of 0.2 mL/min. Mass analyses were performed using a Finnigan Matt LCQ™ ESI-IT mass spectrometer (Thermo Fisher Scientific) in positive ionization mode with a capillary temperature of 160 °C, a capillary voltage and a spray voltage of 3 V and 4.50 kV, respectively. Nitrogen flow rates were 60 and 0 (arbitrary units) respectively for sheath gas and auxiliary gas. The parameters of ion optic transmission were 50 V, -3.25 V, -5.5 V and 400 V for tube lens offset, multipole 1 offset, multipole 2 offset and multipole RF amplifier respectively.

Analyzed under neutral conditions and in positive mode by electrospray ionization-ion trap-mass spectrometry (ESI-IT-MS), long chain peptaibols appear principally as doubly charged sodium adduct ions [M + 2Na]²⁺ and short chain peptaibols appear as singly charged sodium adduct ions [M + Na]⁺. All data acquisitions and reprocessings were performed using LCQ Xcallibur 1.3 software (Thermo Fisher Scientific). For each standard, the analyzed ions were:

- PEP-11: $[M + Na]^{+} = 1197$

- PEP-14: $[M + Na]^{+} = 1417$ and $[M + 2Na]^{2+} = 720$

- PEP-20: $[M + Na]^{+} = 1985$ and $[M + 2Na]^{2+} = 1004$.

Concerning the peptaibols produced by *Trichoderma* sp., the expected diagnostic ions of each subgroup are presented in Table 2.

The short chain peptaibols were quantified using the standard calibration curve performed with PEP-11 and PEP-14 for 11- and 14-residue peptaibols respectively. The long-chain peptaibols
<table><caption>**Table 1**

Trichoderma strain descriptions, origins and putative peptaibol productions.</caption><thead><tr><th>Strain</th><th>Species</th><th>Isolated from</th><th>Geographical origin</th><th>GenBank accession number</th><th>Putative peptaibol subgroups</th></tr></thead><tbody><tr><td>MMS 13</td><td><i>T. harzianum</i></td><td>Cockles</td><td>Le croisic</td><td>JQ653065</td><td>11-, 18-res</td></tr><tr><td>MMS 19</td><td><i>T. citrinoviride</i></td><td>Cockles</td><td>Le croisic</td><td>JQ608473</td><td>11-, 20-res</td></tr><tr><td>MMS 58</td><td><i>T. longibrachiatum</i></td><td>Mussels</td><td>La tara</td><td>JQ653055</td><td>11-, 20-res</td></tr><tr><td>MMS 151</td><td><i>T. longibrachiatum</i></td><td>Mussels</td><td>Tharon</td><td>JQ653057</td><td>11-, 20-res</td></tr><tr><td>MMS 510</td><td><i>T. reesei</i></td><td>Sediment</td><td>La couplasse</td><td>GU947793</td><td>11-, 20-res</td></tr><tr><td>MMS 639</td><td><i>T. atroviride</i></td><td>Seawater</td><td>La couplasse</td><td>JQ653056</td><td>11-, 17-, 19-res</td></tr><tr><td>MMS 755</td><td><i>T. capillare</i></td><td>Sediment</td><td>Le collet</td><td>JQ653068</td><td>11-, 20-res</td></tr><tr><td>MMS 852</td><td><i>T. capillare</i></td><td>Sediment</td><td>Port-Giraud</td><td>JQ653070</td><td>11-, 20-res</td></tr><tr><td>MMS 927</td><td><i>T. atroviride</i></td><td>Sediment</td><td>Le croisic</td><td>GU947796</td><td>17-, 19-res</td></tr><tr><td>MMS 975</td><td><i>T. ghanense</i></td><td>Sediment</td><td>Le croisic</td><td>GU947792</td><td>11-, 20-res</td></tr><tr><td>MMS 1255</td><td><i>T. harzianum</i></td><td>Sediment</td><td>Port du bec</td><td>JQ653081</td><td>11-, 14-res</td></tr><tr><td>MMS 1541</td><td><i>T. pleuroticola</i></td><td>Cockles</td><td>Bonne anse</td><td>JQ653064</td><td>11-, 14-, 18-res</td></tr><tr><td>T22 (ATCC 20847)</td><td><i>T. harzianum</i></td><td>Commercial BCA</td><td>-</td><td></td><td>11-, 14-, 18-res</td></tr></tbody></table>

<table><caption>**Table 2**

Characteristic target [M+Na]$^+$ or [M+2Na]$^{2+}$ ions of peptaibol subgroups.</caption><thead><tr><th>11-res<br>[M+Na]$^+$</th><th>14-res<br>[M+Na]$^+$</th><th>17-res<br>[M+2Na]$^{2+}$</th><th>18-res<br>[M+2Na]$^{2+}$</th><th>19-res<br>[M+2Na]$^{2+}$</th><th>20-res<br>[M+2Na]$^{2+}$</th></tr></thead><tbody><tr><td>1141</td><td>1465</td><td>824</td><td>879</td><td>963</td><td>985</td></tr><tr><td>1155</td><td>1479</td><td>831</td><td>882</td><td>970</td><td>992</td></tr><tr><td>1169</td><td>1493</td><td>832</td><td>886</td><td>977</td><td>999</td></tr><tr><td>1183</td><td>1506</td><td>838</td><td>889</td><td>984</td><td>1005</td></tr><tr><td>1197</td><td>1520</td><td>839</td><td>893</td><td>991</td><td></td></tr><tr><td>1211</td><td>1534</td><td>845</td><td>896</td><td>998</td><td></td></tr><tr><td>1225</td><td></td><td>846</td><td>900</td><td></td><td></td></tr><tr><td></td><td></td><td>853</td><td>903</td><td></td><td></td></tr><tr><td></td><td></td><td></td><td>907</td><td></td><td></td></tr><tr><td></td><td></td><td></td><td>910</td><td></td><td></td></tr><tr><td></td><td></td><td></td><td>914</td><td></td><td></td></tr><tr><td></td><td></td><td></td><td>917</td><td></td><td></td></tr><tr><td></td><td></td><td></td><td>925</td><td></td><td></td></tr></tbody></table>

(17–20-residues) were quantified using the calibration curve performed with PEP-20.

The quantification of peptaibols according to their length in all samples (standards included) were performed in the same conditions and during a single analysis. Nevertheless, the target ions for the quantification of the peptaibols in crude extracts were not the same that the characteristic ions tracked for the standards PEP-14 and PEP-20. On the contrary, the target ion used for the detection of PEP-11 (m/z = 1197) was among the target ions analyzed for the 11-residue peptaibol quantification in crude extracts (Table 2). *Trichoderma* sp. strains potentially produce peptaibols which exhibit the same target ion than the synthesized one, PEP-11, even if the sequences are not the same.

## 2.7. Method validation

The validation process was conducted according to the guidance for industry bioanalytical method validation, recommended by Food and Drug Administration (FDA) [33].

Co-mingled standards (PEP-11, PEP-14 and PEP-20) were prepared at 6 concentration levels: 0.10; 0.25; 0.50; 0.75; 1.00 and 1.25 µg/mL with HPLC quality grade methanol. The calibration curves were constructed by plotting the chromatographic peaks area versus the corresponding concentrations. The data were fitted using a linear regression with 1/X$^2$ weighting factor.

### 2.7.1. Specificity

Specificity of the methods was determined by analyzing a methanol sample (neat solvent used for the sample preparation) and by comparing it to a calibration level in order to investigate the potential interferences at the retention times of the standards and naturally produced peptaibols. If there is no interfering peak at the

expected retention times for the selected ions, the specificity of the methods will be validated.

### 2.7.2. Linearity

The linearity was evaluated from 3 injections of the calibration levels of co-mingled standard solutions. A linear model was applied to the triplicates. For each standard, the slope and the y-intercept of the calibration curve were calculated as well as the coefficient of determination. The linearity was considered as satisfactory when the coefficient of determination (R$^2$) of the calibration curves was higher than 0.999 for each curve [34]. Then, the variance analysis was also performed to validate the linear models in terms of regression and linearity.

### 2.7.3. LOD and LOQ

The limits of detection (LOD) were defined for each compound as the lowest concentration with a signal-to-noise ratio of 3 and was calculated as 3σ/S (σ is the standard deviation of the linear regression and S the slope of the calibration curve) and the limits of quantification (LOQ) were defined as the lowest concentration with a signal-to-noise ratio of 10 and calculated as 10σ/S.

### 2.7.4. Intra-assay precision and accuracy (repeatability)

Three calibration levels (0.50, 0.75 and 1.00 µg/mL) were injected 6 times in a single run. The means, the standard deviations, the coefficients of variation (%CV) and the percent deviations of the back calculated concentrations (%bias) were calculated for each standard level. The coefficients of variation of the six injections of each level were considered as acceptable when inferior to 15%. The back calculated concentrations of each level had to be within 15% deviation from the nominal value.