Bioconcentration and biotransformation of organophosphorus flame retardants (PFRs) in common carp (Cyprinus carpio)

Environment International 126 (2019) 512–522

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Bioconcentration and biotransformation of organophosphorus flame retardants (PFRs) in common carp (Cyprinus carpio)

T

Bin Tanga,b,c, Giulia Pomab, Michiel Bastiaensenb, Shan-Shan Yinb, Xiao-Jun Luoa, , Bi-Xian Maia, ⁎ Adrian Covacib, ⁎

a

State Key Laboratory of Organic Geochemistry, Guangdong Key Laboratory of Environmental Resources Utilization and Protection, Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, Guangzhou 510640, PR China b Toxicological Centre, University of Antwerp, Universiteitsplein 1, 2610 Wilrijk, Belgium c University of Chinese Academy of Sciences, Beijing 100049, PR China

ARTICLE INFO

ABSTRACT

Handling Editor: Yong-Guan Zhu

Understanding the bioaccumulation and biotransformation of xenobiotic compounds is critical for evaluating their fate and potential toxicity in vivo. In the present study, the tissue specific accumulation and depuration of seven organophosphorus flame retardants (PFRs) in common carp (Cyprinus carpio) were investigated after exposing the fish to an environmental relevant level of PFRs. The log Kow of PFRs was significantly negatively correlated to the percentages of individual PFRs to the total PFRs in serum (p < 0.04), whereas significantly positive correlations were observed in all other tissues (p < 0.02). Significant correlations (p < 0.01) between the log Kow of PFRs and their log bioconcentration factor (BCFww) were also found in all investigated tissues except for serum. This suggests that the hydrophobicity of PFRs played a significant role in the distribution and body compartment accumulation of PFRs in common carp. The bioaccumulation potential of PFRs in serum was different from the other tissues, probably due to its specific properties. Dialkyl and/or diaryl phosphate esters (DAP) and hydroxylated PFRs (HO-PFRs) were quantified as the major metabolites. Their levels in liver and intestine were significantly higher than in other tissues. Biotransformation processes also played a crucial role in the accumulation of PFRs in fish. Our results provide critical information for further understanding the bioconcentration, tissue distribution and metabolism of PFRs in fish.

Keywords: Organophosphorus flame retardants Tissue-specific bioconcentration Metabolites Common carp

1. Introduction With the gradual phasing out of brominated flame retardants (BFRs) due to their persistence, bioaccumulation, and toxicity (Covaci et al., 2011; Poma et al., 2018), the production and use of organophosphorus flame retardants (PFRs) as primary substitutes have increased significantly in recent years (van der Veen and de Boer, 2012; Wei et al., 2015). PFRs are mainly used as additive substances, rather than chemically bonded, to the diverse substrate materials such as textile, plastics, foams, lubricants, and paints (van der Veen and de Boer, 2012; Wei et al., 2015). This can result in their release into the environment during usage and disposal of commercial products. Consequently, PFRs are now ubiquitously present in various environmental compartments, including indoor air and dust (Abdallah and Covaci, 2014; Ali et al., 2012; Xu et al., 2016; Zheng et al., 2015), wastewater and sludge (Zeng et al., 2014), surface and drinking water (Ding et al., 2015; Rodil et al., 2012), sediment and soil (Cao et al., 2017; Cao et al., 2012; Tan et al.,

⁎

2016; Wan et al., 2016), biological samples, including human milk and placenta (Ding et al., 2016; Kim et al., 2014), and aquatic biota (Kim et al., 2011; Malarvannan et al., 2015; Zhao et al., 2018). The increasing concentrations of PFRs found in the environment, biota, and human samples raise health concerns because of their potential toxic effect (Wei et al., 2015). Some PFRs, such as tris(2-chloroisopropyl) phosphate (TCIPP), tris(2-chloroethyl) phosphate (TCEP), tris(1,3-dichloro-2-propyl) phosphate (TDCIPP), tri-n-butyl phosphate (TNBP), tris(2-butoxyethyl) phosphate (TBOEP), and triphenyl phosphate (TPHP) are suspected to be carcinogenic, mutagenic, embryotoxic or neurotoxic (van der Veen and de Boer, 2012; Wang et al., 2015; Wei et al., 2015). Chronic exposure to 2-ethylhexyl diphenyl phosphate (EHDPHP) could also be worrisome due to its potential for developmental toxicity (UK Environment Agency, 2009). More specifically, toxicological effects have also been found in fish after exposure to PFRs at environmentally relevant levels. For instance, TPHP showed great cardiotoxicity in zebrafish (Danio rerio) embryos at

Corresponding authors. E-mail addresses: [email protected] (X.-J. Luo), [email protected] (A. Covaci).

https://doi.org/10.1016/j.envint.2019.02.063 Received 6 January 2019; Received in revised form 22 February 2019; Accepted 25 February 2019 0160-4120/ © 2019 The Authors. Published by Elsevier Ltd. This is an open access article under the CC BY license (http://creativecommons.org/licenses/BY/4.0/).

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an exposure level of 100 μg L−1 (Du et al., 2015). A significant decrease in fecundity of zebrafish was observed after exposure to 40 and 200 μg L−1 of TPHP or TDCIPP for 21 days (Liu et al., 2013). TDCIPP exposure in the early embryo stage of zebrafish induced genome-wide hypomethylation (Volz et al., 2016), which can be transferred to the offspring of zebrafish, causing thyroid endocrine disruption and developmental neurotoxicity after long-term exposure at environmentally relevant levels (20 and 100 μg L−1) (Wang et al., 2015). Bioaccumulation potential is a crucial criterion for assessing the ecological risks of PFRs (Wang et al., 2017b; Wei et al., 2015). However, only few studies have investigated the bioaccumulation potential and bioconcentration kinetics of PFRs in fish (McGoldrick et al., 2014; Wang et al., 2017b; Zhao et al., 2018). The bioconcentration factors (BCFs) of TCEP (0.7−2.2), TDCIPP (3−113), tributyl phosphate (TBP, the mixture of TNBP and tris(isobutyl) phosphate (TIBP), 6−35), and TPHP (32−500) have been evaluated in goldfish (Carassius auratus), killifish (Oryzias latipes), and zebrafish in laboratory conditions (Sasaki et al., 1981; Wang et al., 2016), and the BCFs of eight PFRs were measured in topmouth gudgeon (Pseudorasbora parva), crucian carp (Carassius auratus) and loach (Misgurnus anguillicaudatus) from locations around Beijing, China, with values ranging from 27.8 for TDCIPP to 1983 for tris(2-ethylhexyl) phosphate (TEHP) (Hou et al., 2017). A laboratory exposure study on zebrafish showed that the BCFww (BCF on a basis of wet weight) of seven PFRs ranges from 0.5 to 364, with TPHP and tri-p-cresyl phosphate (TCP) exhibiting higher BCFww than the others (Wang et al., 2017b). Several in vivo and in vitro studies have demonstrated that PFRs can be metabolized to diesters (dialkyl and diaryl phosphate esters, DAPs) and hydroxylated PFRs (HO-PFRs) through phase-I biotransformation mechanisms, and further to glucuronide- and sulfate-conjugates via phase-II biotransformation mechanisms in human or chicken (Su et al., 2015; Van den Eede et al., 2013). Limited studies have reported on the in vivo distribution of DAP metabolites in fish (Wang et al., 2017a; Wang et al., 2017b; Wang et al., 2015), while data on HO-PFRs remain scarce. In a recent in vitro study using fish liver microsomes, HO-PFR metabolites [i.e. 2-hydroxyethyl bis(2-butoxyethyl) phosphate (BBOEHEP), and bis (2-butoxyethyl) 3′-hydroxy-2-butoxyethyl phosphate (3-OH-TBOEP) for TBOEP, and dibutyl-3-hydroxybutyl phosphate (3-OH-TNBP) for TNBP] were found as the predominant metabolites compared to the DAP metabolites [bis(2-butoxyethyl) phosphate (BBOEP) and di-n-butyl phosphate (DNBP) for TBOEP and TNBP, respectively] (Hou et al., 2018). This indicated why it is significant to study the presence of HO-PFRs and DAPs in an in vivo study in fish. The biotransformation products could be used as biomarkers to assess the exposure to their parent triester PFRs in biota samples, but also to investigate the in vivo accumulation and biotransformation of PFRs. Even though this information is essential for evaluating their toxicity, the fate of PFRs in fish remains unknown (Wang et al., 2017a; Wang et al., 2017b). In the present study, common carp (Cyprinus carpio) was exposed via water to seven PFRs (TCEP, TCIPP, TBOEP, TDCIPP, TNBP, TPHP, and EHDPHP), which are frequently detected in aquatic ecosystems. The exposure occurred at an environmental relevant level. The accumulation and distribution of PFRs as well as their major metabolites were investigated in different tissues (i.e. muscle, liver, gonad, gill, intestine, brain, kidney, and serum). The effects of biotransformation on bioconcentration of PFRs in fish tissues were evaluated. The primary objective of this study was to provide critical information on the tissue-specific bioconcentration, depuration, and biotransformation potential of PFRs in fish.

2.2. Fish exposure and sampling Fifty-four juvenile common carps (initial average lengths and weights of 9.71 ± 0.18 cm and 28.64 ± 0.88 g, respectively) with mixed-gender were purchased from a local aquarium market in Guangzhou, China. They were used in the exposure experiment performed following the guidance of OECD No. 305 (OECD, 2012), only with minor modifications. At the beginning of the experiment, six fish were randomly selected for background level analysis. The remaining ones (n = 48) were randomly distributed between two 100-L rectangular glass aquariums. Two thirds of the fish (n = 32) were randomly selected as the treated group, with exposure concentrations of 10 μg L−1 per PFR compound in a mixture of seven (TCEP, TNBP, TBOEP, TCIPP, TDCIPP, TPHP, and EHDPHP). This designed exposure concentration is comparable to the reported environmental concentrations of TCEP, TCIPP, TDCIPP, TPHP and TBOEP in water (Wei et al., 2015). A stock solution containing seven PFRs was prepared in dimethyl sulfoxide (DMSO), and the final proportion of DMSO in the exposure medium was approximately 0.01% (v/v). The remaining 16 specimens were selected as the control group (n = 16) and were exposed only to a solution 0.01% DMSO. Only one exposure concentration was conducted in the present study, since comparable bioaccumulation potential of PFRs in fish tissues were found between the low and high exposure level groups (Wang et al., 2017b), which suggests that bioaccumulation and tissue distribution of PFRs in fish would be independent of the exposure concentration. Previous studies exposed a single PFR compound (Wang et al., 2016) or several PFRs in a mixture with varying concentrations for each PFR compound (Wang et al., 2017b) to the fish. In the present study, the exposure to all seven PFRs occurred at the same concentration with an environmental relevant level (10 μg L−1 per compound), which would ensure the reflection of the actual bioaccumulation potential for PFR compounds with different chemical structures and properties in fish. Each aquarium was filled with filtered dechlorinated tap water (maintained at constant temperature of 24–25 °C, pH 6.5–7.5, 7.8–8.4 mg L−1 of dissolved oxygen, and with 12/12 h light-dark cycle). Fish were acclimatized for two weeks prior to exposure, daily fed with commercial food pellets (Yangzhou Five-Star Aquatic Products Co., LTD, PR China) at 1% rate of their average weight. The feeding was ceased 5 days before the experiment, and then the same feeding rate (1% of fish body weight) as during the acclimation process was kept throughout the experiment. During the exposure period, half of the exposure medium was renewed daily with dechlorinated water freshly spiked with 10 μg L−1 per PFR compound, and the aquariums were kept under the same conditions as during the acclimation process. After the 28-day exposure (i.e. uptake period), the fish were transferred to a new aquarium for a further 14-day depuration. The depuration was conducted in dechlorinated tap water without added PFRs, and the whole medium was renewed daily. Sampling was performed on days 3, 7, 14, 21, and 28 during the uptake period and days 3, 7, and 14 during the depuration period. At each sampling time, four fish were selected randomly from the aquarium. The length and body weight of each fish were measured; blood samples were taken from the dorsal aorta, centrifuged at 3000 rpm for 20 min, and serum was collected. The sera collected on each sampling day were subsequently pooled into one sample. Then, the fish were anesthetized on ice for 20 min and carefully dissected. Seven tissues (gill, liver, gonad, intestine, brain, kidney, and muscle) were collected from each individual fish and the mass of each tissue was recorded. On each sampling day, two fish were randomly sampled from the control group, and treated in the same way as those in the exposure group. Moreover, the fish food pellets were collected on each sampling day and analyzed for PFRs and PFR metabolites as well. Additionally, fish feces were siphoned out of the aquarium shortly after egestion and pooled into eight composite samples (collected on days 1–3, 4–7, 8–14, 15–21, and 22–28, and on days 29–31, 32–35, and

2. Materials and methods 2.1. Chemicals and materials The targeted PFR compounds and their metabolites are listed in Table 1. Details regarding the chemicals and materials used in this study are reported in the Supplementary information (SI). 513

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Table 1 Overview of PFR parent compounds and their respective metabolites considered in this study. Parent compound

Metabolites

DAPa/HO-PFRb

Tri-n-butyl phosphate (TNBP) Tris(chloroethyl) phosphate (TCEP) Tris(2-chloroisopropyl) phosphate (TCIPP)

Di-n-butyl phosphate (DNBP) Bis(chloroethyl) phosphate (BCEP) Bis(1-chloro-2-propyl) phosphate (BCIPP) 1-Hydroxy-2-propyl bis(1-chloro-2-propyl) phosphate (BCIPHIPP) Bis(1,3-dichloro-2-propyl) phosphate (BDCIPP) Diphenyl phosphate (DPHP) 3- and 4-Hydroxyphenyl diphenyl phosphate (HO-TPHP) 4-Hydroxyphenyl phenyl phosphate (4-HO-DPHP) 2-Ethylhexyl phenyl phosphate (EHPHP) 2-Ethyl-5-hydroxyhexyl diphenyl phosphate (5-HO-EHDPHP) Bis(2-butoxyethyl) phosphate (BBOEP) 2-Hydroxyethyl bis(2-butoxyethyl) phosphate (BBOEHEP) Bis(2-butoxyethyl) 3′-hydroxy-2-butoxyethyl phosphate (3-HO-TBOEP)

DAP DAP DAP HO-PFR DAP DAP HO-PFR HO-PFR DAP HO-PFR DAP HO-PFR HO-PFR

Tris(1,3-dichloro-2-propyl) phosphate (TDCIPP) Triphenyl phosphate (TPHP) 2-Ethylhexyldiphenyl phosphate (EHDPHP) Tris(2-butoxyethyl) phosphate (TBOEP)

a b

DAP: dialkyl or diaryl phosphate ester. OH-PFR: hydroxylated PFR.

36–42, respectively pooled). All samples, except for serum, were freezedried, ground into powder, and stored at −20 °C prior to analysis. Finally, 1 L of water was collected from the aquarium on each sampling day and analyzed for the target PFR and PFR metabolites.

2.4. Instrument analysis Chromatographic analysis of PFRs was performed using an Agilent 7890B gas chromatograph, equipped with an Agilent 7693A autosampler with a multimode inlet (MMI), coupled to a 7000C triplequadrupole mass spectrometer (GC–MS/MS, Agilent Technologies Inc., Santa Clara, CA, USA), with an EI source working in electron impact mode following the methodology reported by Poma et al. (2018). Instrumental analysis of PFR metabolites was carried out on an Agilent 1290 Infinity liquid chromatography system coupled to an Agilent 6460 Triple Quadrupole mass spectrometer (LC-MS/MS, Santa Clara, CA, USA) with electrospray ionization (ESI) source according to Bastiaensen et al. (2018). Detailed descriptions of both GC–MS/MS and LC-MS/MS procedures are given in the SI.

2.3. Sample preparation The analytical procedures for the extraction and clean-up of PFRs from fish tissues, food and composite feces samples were performed according to Poma et al. (2018); while extraction and clean-up procedure of PFRs from serum and water were performed according to Loseth et al. (2018) and Ding et al. (2015), respectively. The detailed description of the sample preparation procedures is given in Section S2 of SI. The analytical procedure for extraction of PFR metabolites from water was done according to Been et al. (2017, 2018) and is described in detail in the SI. For the extraction of PFR metabolites from fish tissues, food and feces, ~50 mg of each sample was weighed in a 15-mL glass tube. The sample was added with 1.0 mL methanol (MeOH) and 50 μL internal standards (IS) mixture (TBOEP-d6, TPHP-d15, BBOEHEP-d4, TCEP-d12, BDCIPP-d10, DPHP-d10, BBOEP-d4, and BCEP-d8), vortexed 1 min, ultrasonicated 30 min, and centrifuged at 3500 rpm for 5 min. The supernatant was then transferred to a precleaned glass tube. The extraction procedure was repeated twice, and the supernatant was combined in the same tube. The extracts were gently vortexed, evaporated to near dryness under a gentle stream of nitrogen, and reconstituted with 150 μL ultrapure water (UPW)/MeOH (1:1, v/v). The extracts were then transferred to a 0.2 μm centrifugal filter (VWR) (5 min, 9000 rpm), and finally transferred to labeled amber sample vial for liquid chromatography−tandem mass spectrometry analysis (LC-MS/MS). The analytical procedure for the extraction of PFR metabolites from fish serum was modified from Bastiaensen et al. (2018). Approximately 150–250 μL (162–274 mg) serum was added to a clean glass tube, mixed with 50 μL IS mixture and 1.0 mL of phosphate buffer (1 M, pH 6). Deconjugation enzyme (β-glucuronidase from Supleco E. coli, 25 μL, 2 mg mL−1 in phosphate buffer of pH 6) was added, and incubated for 2 h at 37 °C. Then, 100 μL of formic acid was added to the serum to stop the incubation. The sample was vortexed gently, and then loaded to Bond Elut-C18 cartridge (200 mg, 3 mL, Agilent, Santa Clara, USA), which was pre-conditioned with 3 mL MeOH and 2 mL UPW. The sample tube was washed with 1 mL of UPW, which was then loaded on the cartridge. After extraction, the cartridge was washed with 1.5 mL UPW, followed by eluting the target compounds with 3 mL MeOH. The eluent was carefully evaporated under a gentle nitrogen stream and reconstituted in 150 μL UPW/MeOH (1:1, v/v). The reconstituted sample was filtered using a 0.2 μm centrifugal filter (9000 rpm, 5 min) and then transferred to an amber glass vial for LC-MS/MS analysis.

2.5. Quality assurance and quality control The analytical methods for PFR metabolites in serum and solid matrices (fish tissues, feces and food pellets) were validated by analyses of the native standards in two-level spiked matrices (muscle and serum obtained from the control group spiked at low level 4 ng g−1 and high level 20 ng g−1). Precision, accuracy, IS recoveries, and limits of quantification (LOQs) were calculated. Spiked samples were processed using the described analytical methods, and the measured concentrations were background-corrected by subtracting the average concentrations in the non-spiked samples. For the two spiking levels, PFR metabolites recoveries ranged from 81 to 138% in fish muscle and from 83 to 136% in serum, respectively; the IS recoveries ranged from 92 to 102% in fish muscle and from 80 to 92% in serum, respectively. Details regarding the validation results are presented in Tables S1 and S2. Due to poor method performances, BCEP and its mass-labeled internal standard BCEP-d8 were excluded from the quantification. The quality control check was performed as described in Poma et al. (2018). Two procedural blanks were run in parallel with every batch of samples for the analysis of both PFRs (n = 16) and PFR metabolites (n = 16), to adjust for potential background contamination. Average blank levels per batch were subtracted from the sample results, and a value equal to 3 × SD (standard deviation) of the blank measurement was used as the LOQ before results were reported. LOQs for PFR and PFR metabolites ranged between 0.10 and 1.5 ng g−1 ww for fish tissues, and between 0.10 and 0.72 ng L−1 for water samples. IS recoveries for PFRs and PFR metabolites ranged from 90 ± 8% to 103 ± 11% and from 80 ± 5% to 92 ± 2%, respectively. 2.6. Data analysis The bioconcentration parameters, including uptake rate (k1) and depuration rate constants (k2), half-lives (t1/2), and BCFww of PFRs were 514

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calculated according to equations described in previous studies (OECD, 2012; Wang et al., 2017a, b), and described in details in Section S4 of SI. Statistical analysis was performed with SPSS 21 for Windows (SPSS, Inc., Chicago, IL, USA). The statistical differences in the weights and lengths of common carps, and the levels of PFRs and their metabolites among fish tissues were determined by one-way analysis of variance (ANOVA) with Tukey's post hoc test and paired-samples t-test. Correlations were evaluated using Pearson's correlation. The criterion for significance was set at p < 0.05.

TNBP (15.6–22.1%) > TDCIPP (5.7–16.6%) ≥ TBOEP (4.7–8.2%) > TCIPP (3.0–5.6%) ≥ TCEP (1.2–3.8%) (Fig. 2a). Contrary, the order of the contributions for PFRs in serum was TCIPP (23.6%) > TCEP (21.8%) > TBOEP (17.0%) > TDCIPP (12.6%) > EHDPHP (11.4%) > TPHP (7.5%) > TNBP (6.1%) (Fig. 2a). These trends were further demonstrated by the significant negative correlation between the log Kow and contributions of PFRs in serum (R2 = 0.53, p < 0.04, Fig. 2b), whereas significant positive correlations were found in all other tissues (R2 = 0.66–0.81, p < 0.02, Fig. 2b). This result implied that the hydrophobicity of PFRs played a crucial role in the accumulation of PFRs in common carp. The higher polarity of serum compared to other tissues (Zeng et al., 2013) may result in the preferential accumulation of PFRs with lower log Kow values, i.e. TCEP and TCIPP (log Kow = 1.44 and 2.59, respectively). The levels of PFRs in fish tissues rapidly decreased after three days of depuration period: concentrations of PFRs were one or two orders of magnitude lower than those in the exposure period (Fig. 1). For this reason, the calculation of bioconcentration parameters of PFRs in fish was based on two sampling points (on 28 d and 31 d). In the present study, the estimated k1 for TCEP, TCIPP, TBOEP, TDCIPP, TNBP, TPHP and EHDPHP was respectively in the range of 1.3–7.1, 3.2–7.7, 4.7–23.5, 6.8–62.1, 1.7–59.2, 3.4–85.4 and 4.5–111.0 d−1. For TDCIPP, TNBP, TPHP and EHDPHP, k1 were higher than those of other PFRs in fish tissues, except for serum (Table 2). Comparable values for t1/2 of PFRs were found in the different tissues, and the t1/2 for TCEP, TCIPP, TBOEP, TDCIPP, TNBP, TPHP and EHDPHP was respectively in the range of 9.2–18.3, 10.5–17.0, 13.0–14.5, 9.4–19.8, 8.8–20.2, 9.7–18.6 and 8.8–17.6 h (Table 2). Moreover, TCEP and TCIPP showed higher k1 in serum than those in the other tissues; whereas for the other PFRs (TNBP, TDCIPP, TPHP, EHDPHP and TBOEP), a relatively low k1 was found in serum. This could be related to the low log Kow values for TCEP (1.44) and TCIPP (2.59), which thus tend to remain in the blood (serum) rather than to distribute to other organs. This result was in agreement with the higher contribution of TCEP and TCIPP in serum compared to other tissues. The relatively low t1/2 of PFRs found in tissues of common carp in the present study were consistent with those in the previously studies (van der Veen and de Boer, 2012; Wang et al., 2017b), suggesting that PFRs might be not persistent in fish (van der Veen and de Boer, 2012; Wang et al., 2017b). The BCF is a comprehensive parameter for the different processes of uptake, depuration, and biotransformation. In the present study, the calculated BCFk (kinetic BCF) were close to the BCFss (steady-state BCF) for PFRs (Table 2), while only BCFss (expressed as BCFww hereinafter to compare with those in the previously studies) were used for data analysis. The BCFww of TCEP, TCIPP, and TBOEP in all fish tissues, except for serum, were lower than those of the other PFRs, ranging from 1.0 ± 0.1 to 14.8 ± 0.2, indicating that these three PFRs were relatively less bioaccumulative in fish. The BCFww for TNBP, TDCIPP, TPHP, and EHDPHP respectively ranged from 16.7 ± 0.4 to 31.4 ± 0.7, from 7.5 ± 0.3 to 34.6 ± 1.1, from 21.9 ± 0.8 to 60.5 ± 2.6, and from 25.9 ± 0.7 to 78.5 ± 2.4 in common carp tissues, except for serum. The BCFww for PFRs in fish serum varied in the following order: TCIPP (5.2 ± 0.4) > TDCIPP (4.7 ± 0.6) > TBOEP (4.7 ± 0.4) > TCEP (3.8 ± 0.2) > EHDPHP (3.5 ± 0.1) > TPHP (2.3 ± 0.2) > TNBP (1.7 ± 0.1). Moreover, significant positive correlations between the log Kow of PFRs and their log BCFww in all selected tissues (R2 = 0.87–0.95, p < 0.01; Fig. 3a), with exception of serum, further suggested that the hydrophobicity of PFRs played a significant role in the distribution and body compartment accumulation of PFRs in common carp. Similar results were found in zebrafish in a previous study (Wang et al., 2017b). Meanwhile, the correlations between the BCFww and the lipid contents of fish tissues were examined to investigate the effect of lipid contents on the bioconcentration of PFRs in fish tissues. The BCFww of TDCIPP, TNBP, TPHP, and EHDPHP in common carp tissues exhibited significant correlations with lipid contents (R2 = 0.74–0.99, p < 0.01), while only a significant but weak correlation with the lipid contents of tissues was found for TBOEP

3. Results and discussion 3.1. Background levels Throughout the 42 days of experiment period, no significantly differences were found for the weights and lengths of common carps in both the exposure and control groups (one-way ANOVA, p > 0.05). Therefore, the growth dilution for the concentration of PFRs and PFR metabolites in fish tissues was ignored in this study. The average concentration of each PFR compound in the water during the exposure period was slightly lower than the nominal concentration (10.0 μg L−1) and ranged from 4.3 ± 0.3 μg L−1 (TDCIPP) to 9.1 ± 0.2 μg L−1 (TCEP) (Table 2 and Fig. S1), likely caused by adsorption on glassware. PFR levels in water samples collected during the depuration period and from the control group aquarium were a factor 1000 lower than for the exposure levels and ranged from < LOQ to 5.2 ng L−1. The average concentrations of the seven PFRs ranged from 0.12 ± 0.20 ng g−1 ww (TCEP) to 0.37 ± 0.06 ng g−1 ww (TPHP) in the control fish tissues, from 0.09 ng g−1 dry weight (dw, EHDPHP) to 0.30 ng g−1 dw (TCIPP) in fish food sample, and from 0.08 ng g−1 dw (TDCIPP) to 0.42 ng g−1 dw (TCEP) in fish feces of the control group. Being several orders of magnitude lower than those in the exposed fish, the background levels of parent PFRs in fish are negligible. The concentrations of PFR metabolites were all below LOQ in food, water, fish tissue, and feces samples of the control group. 3.2. Bioconcentration and depuration of PFRs in fish tissues The uptake and depuration kinetic constants of PFRs in the muscle, liver, gonad, brain, gill, kidney, intestine, and serum of common carp were calculated. All targeted PFRs were detected in the selected tissues of common carp after 3 days of exposure. Concentrations of all targeted PFRs in the fish tissues are presented in Fig. 1. A steady-state was achieved for each PFR compound between 3 and 14 d of the exposure period. Specifically, concentrations of all PFRs in serum and gonad achieved steady state in 3 days (Fig. 1). The accumulation of each PFR compound in fish appeared to be tissue-specific, and the concentrations of PFRs at steady-state varied among tissues (Fig. 1). Muscle had relatively low levels for all PFRs; serum contained the lowest levels of PFRs, except for TCEP and TCIPP; liver showed the highest levels for TNBP, TCEP and TBOEP, and intestine contained the highest level of TPHP and EHDPHP, whereas TDCIPP preferentially accumulated in the kidney (Fig. 1). In the present study, all PFR compounds in the exposure medium were set at the same concentration (i.e. 10.0 μg L−1, although the actual concentration of each PFR in water was differed from the nominal concentration due to adsorption and/or other factors, most of the PFR compounds showed similar concentrations in water, Table 2), as thus, the compositions of PFRs in fish tissues can be used for assessing their accumulation potential in these tissues. The relative composition of PFRs in each tissue at the steady-state is presented in Fig. 2a. No statistically significant difference was found for the PFR compositions among fish tissues (oneway ANOVA, p > 0.05), except for serum (in which the PFR compositions was similar to that in the water, Fig. 2a). In all tissues except for serum, a same decreasing order for PFR compositions was found: EHDPHP (28.2–38.1%) > TPHP (22.1–28.7%) > 515

516

Muscle Liver Gonad Brain Gill Kidney Intestine Serum Muscle Liver Gonad Brain Gill Kidney Intestine Serum Muscle Liver Gonad Brain Gill Kidney Intestine Serum Muscle Liver Gonad Brain Gill Kidney Intestine Serum

Tissue

> 1.5 > 7.1 > 2.8 > 3.7 > 2.8 > 1.7 > 1.3 > 6.4 > 3.2 > 5.2 > 4.3 > 5.4 > 7.7 > 6.7 > 6.2 > 7.7 > 7.6 > 23.5 > 14.0 > 15.7 > 8.8 > 10.2 > 11.0 > 4.7 > 9.4 > 20.0 > 32.1 > 19.2 > 17.9 > 62.1 > 27.6 > 6.8

(d−1)

k1

> 1.5 > 1.7 > 1.6 > 1.4 > 1.8 > 1.1 > 0.9 > 1.7 > 1.3 > 1.2 > 1.2 > 1.2 > 1.6 > 1.3 > 1.2 > 1.5 > 1.2 > 1.6 > 1.4 > 1.6 > 1.2 > 1.1 > 1.2 > 1.0 > 1.5 > 1.4 > 1.6 > 0.8 > 1.6 > 1.8 > 1.4 > 1.4

(d−1)

k2

< 11.3 < 9.9 < 10.3 < 12.0 < 9.2 < 15.1 < 18.3 < 10.0 < 13.0 < 13.4 < 14.5 < 14.4 < 10.5 < 13.2 < 13.9 < 11.0 < 13.7 < 10.5 < 11.8 < 10.6 < 14.3 < 14.7 < 13.5 < 17.0 < 11.2 < 11.6 < 10.5 < 19.8 < 10.6 < 9.4 < 12.1 < 11.7

(h)

t1/2

1 4.2 1.7 2.6 1.6 1.6 1.5 3.8 2.5 4.2 3.7 4.7 4.9 5.4 5.2 5.1 6.3 14.8 9.9 10 7.5 9.1 8.9 4.8 6.3 13.9 20.3 22.9 11.4 35 20.1 4. 8

(L kg−1)

BCFk

1.0 ± 0.1 4.3 ± 0.2 1.8 ± 0.1 2.6 ± 0.2 1.6 ± 0.1 1.6 ± 0.1 1.5 ± 0.1 3.8 ± 0.2 2.7 ± 0.1 4.2 ± 0.1 3.8 ± 0.2 4.7 ± 0.5 4.8 ± 0.1 5.6 ± 0.1 4.9 ± 0.1 5.2 ± 0.4 6.3 ± 0.2 14.8 ± 0.2 9.9 ± 0.6 10.2 ± 0.4 7.5 ± 0.6 9.1 ± 0.2 8.9 ± 0.5 4.7 ± 0.4 7.5 ± 0.3 13.8 ± 0.3 20.3 ± 0.1 22.8 ± 0.4 11.4 ± 1.4 34.6 ± 1.1 21.2 ± 0.2 4.7 ± 0.6

(L kg−1)

BCFss

EHDPHP (5.37)

TPHP (4.59)

TNBP (4.0)

(log Kow)

Compound

5.3 ± 0.2

5.2 ± 0.3

5.9 ± 0.2

(μg L−1)

Actual conc.

Muscle Liver Gonad Brain Gill Kidney Intestine Serum Muscle Liver Gonad Brain Gill Kidney Intestine Serum Muscle Liver Gonad Brain Gill Kidney Intestine Serum

Tissue

> 23.4 > 59.2 > 44.2 > 21.3 > 24.3 > 42.0 > 38.7 > 1.7 > 25.9 > 85.4 > 81.8 > 28.1 > 44.6 > 73.3 > 77.9 > 3.4 > 29.5 > 106.6 > 111.0 > 36. 9 > 56.4 > 85.1 > 103.7 > 4.5

(d−1)

k1

> 1.4 > 1.9 > 1.5 > 0.8 > 1.5 > 1.7 > 1.3 > 1.1 > 1.3 > 1.6 > 1.7 > 0.9 > 1.2 > 1.7 > 1.3 > 1.5 > 1.3 > 1.6 > 1.9 > 1.0 > 1.4 > 1.8 > 1.3 > 1.3

(d−1)

k2

< 11.8 < 8.8 < 11.2 < 20.2 < 11.5 < 9.6 < 12.5 < 15.9 < 12.5 < 10.4 < 9.8 < 18.6 < 13.6 < 9.7 < 13.0 < 11.3 < 13.2 < 10.2 < 8.8 < 17.6 < 12.1 < 9.3 < 12.6 < 12.7

(h)

t1/2

16.7 31.4 29.8 25.8 16.7 24.1 29 1.7 19.4 53.2 48 31.4 36.4 42.6 60.6 2.3 23.5 65.3 59 38.9 41 47.4 78.4 3.5

(L kg−1)

BCFk

16.7 ± 0.4 31.4 ± 0.7 29.8 ± 1.7 25.8 ± 0.2 16.7 ± 1.2 24.1 ± 0.6 29.0 ± 1.2 1.7 ± 0.1 21.9 ± 0.8 54.6 ± 0.5 49.4 ± 2.3 31.7 ± 1.1 36.2 ± 0.7 44.9 ± 1.4 60.5 ± 2.6 2.3 ± 0.2 25.9 ± 0.7 66.9 ± 1.3 61.3 ± 1.3 39.5 ± 0.4 38.4 ± 1.4 50.6 ± 0.7 78.5 ± 2.4 3.5 ± 0.1

(L kg−1)

BCFss

k1 (d−1), uptake rate constant; k2 (d−1), depuration rate constant; BCFk, kinetic bioconcentration factor on wet weight basis; BCFss, bioconcentration factor on wet weight basis on the steady-state; t1/2 (h), depuration half-life. Log Kow values are referred to Wei et al. (2015).

4.3 ± 0.3

7.3 ± 0.3

TCIPP (2.59)

TDCIPP (3.8)

9.1 ± 0.2

TCEP (1.44)

5.4 ± 0.2

(μg L−1)

(log Kow)

TBOEP (3.65)

Actual conc.

Compound

Table 2 Bioconcentration parameters of PFRs in tissues of common carp at steady state.

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140

Muscle

120 100 80 60

450 400 350 300 250 200

Liver

80 60 40 20 0

40 20 0

350

Gonad

300

210

TCEP TCIPP TBOEP TDCIPP TNBP TPHP EHDPHP

Brain

180

250 150

Concentration (ng g-1, ww)

200

120

150 100 80 60 40 20 0

90 60 30 0

350 220

Kidney

300

Gill 200

250

180

200

160

150

100 80 60 40 20 0

40 20 0

450

40

Intestine

400

Serum

35 30

350

25 300

20

200

15

150

10

100

5

50

0

0

-5 3

7

14

21

28

31

35

3

42

7

14

21

28

31

35

42

Tmie (d)

Tmie (d)

Fig. 1. Accumulation and depuration of seven PFRs in different tissues (muscle, liver, gonad, brain, gill, intestine and serum) of common carp through aqueous exposure. Error bar indicates ± standard deviation. The dashed line represents the day before depuration.

(R2 = 0.40, p = 0.04). This indicates that lipid contents in tissues also played significant roles in these PFR distributions (Fig. 3b). However, such correlation was absent for TCEP and TCCIP (Fig. 3c), probably due to their relatively lower log Kow values, and thus lower affinity for lipids. Overall, as discussed above, the bioaccumulation potential of PFRs in serum is different from other tissues, most probably due to the specific function of blood. Generally, the distribution of organic chemicals in tissues is mainly driven by passive diffusion to the lipid compartment, and the blood is in continuous contact and equilibrium with organs and tissues where chemicals are deposited. Therefore, blood (serum) works more as a transfer or transport medium than a matrix for deposition of PFRs. In accordance with the results of the present study,

Wang et al. (2017b) have demonstrated that the partition processes between tissues and blood were the major processes accounting for the accumulation of PFRs in zebrafish using a physiologically-based toxicokinetic (PBTK) model. Despite the widespread occurrence of PFRs in aquatic ecosystems (van der Veen and de Boer, 2012; Wei et al., 2015), data on their bioaccumulation potential in aquatic biota, including fish, are still scarce. In a laboratory study by Wang et al. (2017b), the BCFww of TCEP, TNBP and TDCIPP in zebrafish tissues ranged from 0.5 to 66, which was comparable with those found in the present study. On the contrary, the BCFww for TPHP in zebrafish tissues (ranging from 45 to 224) were higher than those in the present study. In a recent exposure study with common carp (Bekele et al., 2018), the BCFww for TBP, 517

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TCEP

(a)

TCIPP

TBOEP

TDCIPP

TNBP

TPHP

EHDPHP

(b)

40

100

Muscle Liver Gonad Brain Gill Kidney Intestine Serum

Relative Composition Percentages (%)

Relative Composition Percentages (%)

90 80 70 60 50 40 30 20

30

2

R = 0.66-0.81, p < 0.02

20

10

2

10 0

R = 0.53, p < 0.04

0 Muscle

Liver

Gonad

Brain

Gill

Kidney Intestine Serum

1

Water

2

3

4 Log Kow

5

6

Fig. 2. Relative composition percentage (%) of each PFR in fish tissues and water at the steady state (a), and correlations between the relative contributions and the log Kow values of PFRs. Log Kow values are referred to Wei et al. (2015).

TCIPP, TDCIPP, TPHP, TEHP, and TCP ranged from 6.54 to 528.15 in different tissues (muscle, liver, intestine and kidney), which were also higher than those found in the present study. The difference in exposure doses, fish size, and species variability could be some of the reasons for having different BCFww values for PFRs between different studies (Bekele et al., 2018).

metabolites in tissues of fish not only helps understanding the metabolism of PFRs, but also provides insights into their pharmacokinetics and their chemical interactions in the body (Greaves and Letcher, 2014; Wang et al., 2017b). DNBP, BDCIPP and DPHP were measured as the major metabolites of TNBP, TDCIPP and TPHP, respectively, in zebrafish (Wang et al., 2017b), and the levels of BDCIPP and DPHP were approximately 1.2 and 2.0 times higher as those of their respective parent PFR compound, whereas the ratio of DNBP to TNBP was calculated as 0.2 (Wang et al., 2017b). Four DAPs (BBOEP, DNBP, diethylhexyl phosphate (DEHP) and DPHP) were quantified at the same

3.3. PFR metabolites in different fish tissues the

and

distribution

of

PFR

a

Muscle Liver 2 Gonad R = 0.87-0.95, p < 0.01 Brain Gill Kidney Intestine Serum

2.0

Log BCFss (L kg-1,ww)

concentration

1.5

TBOEP TDCIPP TNBP TPHP EHDPHP

80

60

b

R2= 0.40-0.99, p < 0.05 40

BCFss (L kg-1,ww)

Characterizing

1.0

20

0 0

1

2

3

4

5

6

7

8

9

6 TCEP TCIPP

c

5

4

0.5

3

2

R = -0.027

2

0.0 1

1

2

3

4

5

Log Kow

6

0

1

2

3

4

5

6

7

8

9

Lipid contents (%)

Fig. 3. Relationships between bioconcentration factors (BCFs) and log Kow of PFRs (a) and between BCFs and lipid contents (b, c) in tissues of common carp. Error bar indicates ± standard deviation. Lipid contents were determined by a gravimetric technique on sub-samples extracted with Acetone/n-Hexane (1:1, v/v). Log Kow values are referred to Wei et al. (2015). 518

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200

Muscle Liver Gonad Brain Gill Kidney Intestine Serum

DPHP 150

100

50

0 80 36

BDCIPP

BCIPHIPP

30

60

24 40

18 12

20

6 0

0

35 70

DNBP

BBOEHEP

30

60

Concentration (ng g-1, ww)

25 50 20

40

15

20

10 10

5

0

0

50

15 EHPHP

40

HO-TPHP

30

10

20 8 5 4 0

0

20

15 BCIPP

5-HO-EHDPHP

15 10 10 5 5

0

0 5

8 BBOEP

3-HO-TBOEP 4

6

3 4 2 2

1

0

0 3

7

14

21 28 Tmie (d)

31

35

42

3

7

14

21 28 Tmie (d)

31

35

42

Fig. 4. The concentrations (ng g−1) of PFR metabolites in different tissues of common carp through aqueous exposure. Error bar indicates ± standard deviation. The dashed line represents the day before depuration.

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0.6

a) TNBP

DNBP

c) TCIPP

BCIPP BCIPHIPP

0.5 0.4 0.3 0.2

DAPs or OH-PFRs to PFR ratio

0.1 0.0 1.6 1.4

0.8 0.7 0.6 0.5 0.4 0.3 0.2 0.1 0.0 1.2

b) TDCIPP

d) TPHP

1.1

BDCIPP

DPHP HO-TPHP

1.0 1.2 0.6

0.6

0.4

0.4

0.2

0.2

0.0 0.5

0.0 e) TBOEP

0.4

BBOEP 3-HO-TBOEP BBOEHEP

EHPHP 5-HO-EHDPHP

0.2

0.2

0.1

0.1 e r d n l y e scl ive na rai Gilidneestin erum Mu L Go B K Int S

f) EHDPHP

0.3

0.3

0.0

0.4

0.0

e r d n l y e scl ive na rai Gilidneestin erum K Int S Mu L Go B

ter ces Wa Fe

ter ces Wa Fe

Fig. 5. The ratio of the major PFR metabolites to their parent compounds in tissues of common carp. Error bar indicates ± standard deviation.

order of magnitude as their parent PFRs (TBOEP, TNBP, TEHP, and TPHP and EHDPHP) in fish from rivers around Beijing, China (Hou et al., 2017). These results are in accordance with the rapid biotransformation of PFRs in biota (Hou et al., 2016), suggesting that biotransformation is an important factor affecting the accumulation of PFRs in fish. As shown in Fig. 4, all targeted DAPs were detected in fish tissues during the exposure period, and the concentrations were in the range of 1.07 ± 0.39 ng g−1 ww for EHPHP in gill (day 28) to 154.1 ± 10.1 ng g−1 ww for DPHP in liver (day 14). Generally, the levels of DAPs in the liver and intestine were higher than those in other tissues (one-way ANOVA, p < 0.01), indicating that the hepatobiliary system (liver-bile-intestine) plays a key role in the metabolism and excretion of PFRs in fish (Wang et al., 2017b). The concentrations of four DAPs (BBOEP, DNBP, DEHP and DPHP) found in tissues of crucian carp and loach from locations around Beijing, China showed the same order of variation: liver > kidney > intestine > muscle > ovary (Hou et al., 2017). This was in agreement with the results of the present study and with those for DNBP, BDCIPP, and DPHP in zebrafish (Wang et al., 2017b). Of the six targeted HO-PFRs, 4-HO-DPHP was below the limit of quantification (LOQ) in all samples, while BBOEHEP and BCIPHIPP were detected in all tissues, except for kidney; 3-HO-TBOEP and 5-HOEHDPHP were detected in the liver, intestine, serum, and gill; and HOTPHP was detected only in the liver, intestine and serum (Fig. 4). The concentrations of HO-PFRs ranged from 0.80 ± 0.10 ng g−1 ww for 3HO-TBOEP in gill (day 28) to 32.9 ± 0.01 ng g−1 ww for BCIPHIPP in the intestine (day 14) (Fig. 4). Both DAPs and HO-PFRs have already been considered as target biomarkers in human biomonitoring studies (Bastiaensen et al., 2018; Van den Eede et al., 2015b). However, to the best of our knowledge, this is the first study quantifying HO-PFRs to

study the in vivo biotransformation in fish (Fig. 4). The concentrations of BBOEHEP were significantly higher than those of BBOEP in all tissues (t-test, p < 0.01), except for kidney (in which BBOEHEP was below the LOQ), suggested HO-PFRs was the dominant metabolite of TBOEP in fish, which was consistent with the results found in in vitro study in fish (Hou et al., 2018) and humans (Bastiaensen et al., 2018; Van den Eede et al., 2015a). Significantly higher concentrations of BCIPHIPP than BCIPP as metabolites of TCIPP were also observed in liver, brain, gill, intestine, and serum (t-test, p < 0.05). The elimination of PFR metabolites from tissues was fast since DAPs and HO-PFRs could not be detected in fish tissues after three days of depuration (Fig. 4). The formed metabolites are generally more hydrophilic than the PFR parent compounds, and thus are cleared faster from the body and released into the water. Actually, all metabolites detected in fish tissues were found in water (ranging from 10.0 ng L−1 for BBOEP in day 7 to 2594 ng L−1 for BDCIPP in day 7) and in feces (ranging from 2.4 ng g−1 dw for 3-HO-TBOEP in day 3 to 136.5 ng g−1 dw for DPHP in day 14) during the exposure period. Furthermore, the PFR metabolites were < LOQ in water during the depuration period (Fig. S1). The high levels of DAPs and HO-PFR metabolites detected in the present study in the fish tissues and in water suggest the importance of monitoring these metabolites in aquatic ecosystems. Moreover, the levels of DAPs and HO-PFRs in liver, intestine, and feces were significantly correlated to those in water (R2 = 0.51–0.64, p < 0.01, Fig. S2a, b and c), further indicating the easy release of the metabolism products from the liver to the intestine and subsequently into the surrounding water. Besides, a weak but significant correlation (R2 = 0.16, p = 0.011, Fig. S2d) between the concentration of PFR metabolites in serum and those in water was found as well, implying the possibility of elimination of PFR metabolites through gill via the transportation of blood flow (Wang et al., 2016). 520

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3.4. Evaluation of biotransformation effects on the bioconcentration of PFRs in fish

research scholarship (Nos. 201704910738 and 201706320119, respectively) provided by the China Scholarship Council for their research stays at the University of Antwerp. Michiel Bastiaensen and Dr. Giulia Poma acknowledge the provision of their fellowships from the University of Antwerp. This is contribution No.IS-2665 and SKLOGA201602 from GIGCAS.

Biotransformation processes could strongly impact the extent to which hydrophobic organic chemicals accumulate in fish (Ashauer et al., 2012; Wang et al., 2016; Wang et al., 2017b). The metabolite/ parent ratio (MPR) can provide a convenient way to characterize the bioaccumulation potential of the metabolite (Hou et al., 2017). In the present study, the MPRs of PFR metabolites (DAPs and HO-PFRs) to their parent compounds were calculated in fish tissues, in feces and water at steady-state (Fig. 5). The MPR values for each PFR pair varied among tissues and ranged from 0.07 ± 0.01 for TNBP in gill to 1.46 ± 0.16 for TCIPP in the intestine (Fig. 5). As the liver is the main organ for the metabolism of xenobiotic chemicals, the MPR for the main metabolites of each PFR in the liver could be used as a rough evaluation for the biotransformation effect on the bioconcentration of PFRs in fish. The MPR values for PFRs in fish liver ranged from 0.11 ± 0.02 for EHDPHP to 1.36 ± 0.15 for TCIPP. Although PFR metabolites are likely to be more hydrophilic, and thus more quickly cleared from the liver and less accumulated than the parent compounds, the quantification of PFR metabolites at similar levels as their parent compounds found in the present study indicate an intensive transformation of PFRs, and a consequent substantial reduction of accumulation in the fish body. Additionally, significant negative correlations were observed between the MPR values of PFRs and their log BCFww in fish muscle, liver, gill and intestine (R2 = 0.59–0.76, p < 0.05; Fig. S3), further indicate the strongly effect of biotransformation processes on the bioconcentration of PFRs in fish. However, it is worth noting that the phase-I metabolites of PFRs quantified in the present study can further be transformed into glucuronide- and sulfate-conjugates via phase-II biotransformation. As such, the lack of quantification of such phase-II metabolites would result in the underestimation of the biotransformation effect on the bioconcentration of PFRs in the present study. A recent in vivo exposure study qualitatively demonstrated that DAPs accounted for > 85% of the peak area of TPHP, TDCIPP, TCEP, TBOEP and TNBP metabolites in the zebrafish liver, while HO-PFRs have not been detected (Wang et al., 2017a). This could be attributed to the rapid conjugation with glucuronic acid or other phase-II reaction of HO-PFRs occurring in the body (Wang et al., 2017a).

Appendix A. Supplementary data Supplementary data to this article can be found online at https:// doi.org/10.1016/j.envint.2019.02.063. References Abdallah, M.A., Covaci, A., 2014. Organophosphate flame retardants in indoor dust from Egypt: implications for human exposure. Environ. Sci. Technol. 48, 4782–4789. Ali, N., Dirtu, A.C., Van den Eede, N., Goosey, E., Harrad, S., Neels, H., Mannetje, A., Coakley, J., Douwes, J., Covaci, A., 2012. Occurrence of alternative flame retardants in indoor dust from New Zealand: indoor sources and human exposure assessment. Chemosphere 88, 1276–1282. Ashauer, R., Hintermeister, A., O'Connor, I., Elumelu, M., Hollender, J., Escher, B.I., 2012. Significance of xenobiotic metabolism for bioaccumulation kinetics of organic chemicals in Gammarus pulex. Environ. Sci. Technol. 46, 3498–3508. Bastiaensen, M., Xu, F., Been, F., Van den Eede, N., Covaci, A., 2018. Simultaneous determination of 14 urinary biomarkers of exposure to organophosphate flame retardants and plasticizers by LC−MS/MS. Anal. Bioanal. Chem. 410, 7871–7880. Been, F., Bastiaensen, M., Lai, F.Y., van Nuijs, A.L.N., Covaci, A., 2017. Liquid chromatography-tandem mass spectrometry analysis of biomarkers of exposure to phosphorus flame retardants in wastewater to monitor community-wide exposure. Anal. Chem. 89, 10045–10053. Been, F., Bastiaensen, M., Lai, F.Y., Libousi, K., Thomaidis, N.S., Benaglia, L., Esseiva, P., Delemont, O., van Nuijs, A.L.N., Covaci, A., 2018. Mining the chemical information on urban wastewater: monitoring human exposure to phosphorus flame retardants and plasticizers. Environ. Sci. Technol. 52, 6996–7005. Bekele, T.G., Zhao, H., Wang, Y., Jiang, J., Tan, F., 2018. Measurement and prediction of bioconcentration factors of organophosphate flame retardants in common carp (Cyprinus carpio). Ecotoxicol. Environ. Saf. 166, 270–276. Cao, D., Guo, J., Wang, Y., Li, Z., Liang, K., Corcoran, M.B., Hosseini, S., Bonina, S.M., Rockne, K.J., Sturchio, N.C., Giesy, J.P., Liu, J., Li, A., Jiang, G., 2017. Organophosphate esters in sediment of the Great Lakes. Environ. Sci. Technol. 51, 1441–1449. Cao, S., Zeng, X., Song, H., Li, H., Yu, Z., Sheng, G., Fu, J., 2012. Levels and distributions of organophosphate flame retardants and plasticizers in sediment from Taihu Lake, China. Environ. Toxicol. Chem. 31, 1478–1484. Covaci, A., Harrad, S., Abdallah, M.A., Ali, N., Law, R.J., Herzke, D., de Wit, C.A., 2011. Novel brominated flame retardants: a review of their analysis, environmental fate and behaviour. Environ. Int. 37, 532–556. van der Veen, I., de Boer, J., 2012. Phosphorus flame retardants: properties, production, environmental occurrence, toxicity and analysis. Chemosphere 88, 1119–1153. Ding, J., Shen, X., Liu, W., Covaci, A., Yang, F., 2015. Occurrence and risk assessment of organophosphate esters in drinking water from eastern China. Sci. Total Environ. 538, 959–965. Ding, J., Xu, Z.M., Huang, W., Feng, L., Yang, F., 2016. Organophosphate ester flame retardants and plasticizers in human placenta in eastern China. Sci. Total Environ. 554−555, 211–217. Du, Z., Wang, G., Gao, S., Wang, Z., 2015. Aryl organophosphate flame retardants induced cardiotoxicity during zebrafish embryogenesis: by disturbing expression of the transcriptional regulators. Aquat. Toxicol. 161, 25–32. Greaves, A.K., Letcher, R.J., 2014. Comparative body compartment composition and in ovo transfer of organophosphate flame retardants in North American Great Lakes herring gulls. Environ. Sci. Technol. 48, 7942–7950. Hou, R., Xu, Y., Wang, Z., 2016. Review of OPFRs in animals and humans: absorption, bioaccumulation, metabolism, and internal exposure research. Chemosphere 153, 78–90. Hou, R., Liu, C., Gao, X., Xu, Y., Zha, J., Wang, Z., 2017. Accumulation and distribution of organophosphate flame retardants (PFRs) and their di-alkyl phosphates (DAPs) metabolites in different freshwater fish from locations around Beijing, China. Environ. Pollut. 229, 548–556. Hou, R., Huang, C., Rao, K., Xu, Y., Wang, Z., 2018. Characterized in vitro metabolism kinetics of alkyl organophosphate esters in fish liver and intestinal microsomes. Environ. Sci. Technol. 52, 3202–3210. Kim, J.W., Isobe, T., Chang, K.H., Amano, A., Maneja, R.H., Zamora, P.B., Siringan, F.P., Tanabe, S., 2011. Levels and distribution of organophosphorus flame retardants and plasticizers in fishes from Manila Bay, the Philippines. Environ. Pollut. 159, 3653–3659. Kim, J.W., Isobe, T., Muto, M., Tue, N.M., Katsura, K., Malarvannan, G., Sudaryanto, A., Chang, K.H., Prudente, M., Viet, P.H., Takahashi, S., Tanabe, S., 2014. Organophosphorus flame retardants (PFRs) in human breast milk from several Asian countries. Chemosphere 116, 91–97. Liu, X., Ji, K., Jo, A., Moon, H.B., Choi, K., 2013. Effects of TDCPP or TPP on gene transcriptions and hormones of HPG axis, and their consequences on reproduction in

4. Conclusions In the present study, the accumulation and depuration of seven major PFRs in different tissues of common carp were investigated. Significant positive correlations were observed between the log Kow of PFRs and their log BCFww in all investigated tissues, except for serum. The accumulation of PFRs in serum was different from the other tissues. In addition to the hydrophobicity of PFRs, the lipid contents in fish tissues also play a crucial role in the accumulation of PFRs in fish. The DAPs and HO-PFRs metabolites quantified in fish tissues demonstrated an intensive biotransformation of PFRs, and a consequent substantial reduction of accumulation in fish. Additionally, more effort should be made regarding the toxicological effect assessment and environmental monitoring of PFR metabolites in the aquatic ecosystem, given their high levels detected in the tissues of exposed fish and in water in the present study. Acknowledgements This work was financially supported by the National Natural Science Foundation of China (Nos. 41877386 and 41673100), the University of Antwerp, the INTERWASTE project (grant agreement 734522) funded by the European Commission (Horizon 2020), and the Key Research Program of Frontier Sciences, the Chinese Academy of Sciences (QYZDJ-SSW-DQC018). Bin Tang and Shan-Shan Yin acknowledge 521

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