(Carassius auratus) exposed to perfluorooctanoic acid

Science of the Total Environment 653 (2019) 1395–1406

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Molecular and phenotypic responses of male crucian carp (Carassius auratus) exposed to perfluorooctanoic acid Huike Dong a, Guanghua Lu a,b,⁎, Zhenhua Yan a, Jianchao Liu a, Yong Ji c a b c

Key Laboratory of Integrated Regulation and Resources Development of Shallow Lakes of Ministry of Education, College of Environment, Hohai University, Nanjing 210098, China Water Conservancy Project & Civil Engineering College, Tibet Agriculture & Animal Husbandry University, Linzhi 860000, China College of Water Conservancy and Ecological Engineering, Nanchang Institute of Technology, Nanchang 330099, China

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• PFOA preferred to bioconcentrate in protein-rich tissues of fish via flowthrough exposure. • PFOA could induce neurotoxicity and disrupt lipid metabolism at different molecular levels. • PFOA might cause peroxisome proliferation in fish and thus mitochondrial dysfunction. • Phenotypic changes were supposed to be results of related biochemical alterations.

a r t i c l e

i n f o

Article history: Received 6 August 2018 Received in revised form 1 November 2018 Accepted 2 November 2018 Available online 07 November 2018 Editor: Daniel Wunderlin Keywords: PFOA Bioconcentration Biochemical responses Histology Behaviour Carassius auratus

a b s t r a c t Perfluorooctanoic acid (PFOA) has long been produced and widely used due to its excellent water and oil repellent properties. However, this trend has facilitated to the ubiquitous existence of PFOA in environmental matrix, and the potential ecotoxicity on aquatic organisms has not been fully elucidated. To study the tissue-specific bioconcentration and the nervous system- and energy-related biochemical effects of PFOA, as well as the phenotypic alterations by this chemical, male crucian carp (Carassius auratus) were exposed to gradient concentrations of PFOA (nominal 0.2, 10, 500 and 25,000 μg/L) in a flow-through apparatus for 7 days. PFOA was enriched in tissues following an order of blood N kidney ≥ liver N gill N brain N muscle. The bioconcentration factors ranged from 0.1 to 60.4. Acetylcholinesterase activity in the fish brain was inhibited, while liver carboxylesterase was induced in most cases and attenuated with time. The acyl-CoA oxidase activity was dose-dependently elevated and accompanied by a decline of ATP contents. PFOA treatments also inhibited the activity of the electron transport system (ETS). At the transcriptional level, ETS component complexes II and IV were concordantly depressed, and ATP synthesis was also downregulated. The mRNA level of peroxisome proliferator activated receptor α was increasingly upregulated, with related downstream genes upregulated in varying degrees. The phenotypes showed patterns of increased liver pathology and reduced swimming activity. In summary, PFOA leads to adverse effects in Carassius auratus related to multiple aspects, which may be associated with the nervous system, fundamental energy metabolism and other unpredictable factors. The results obtained in this study are expected to help clarify the PFOA toxic mechanisms on energy relevance. © 2018 Elsevier B.V. All rights reserved.

⁎ Corresponding author at: Key Laboratory of Integrated Regulation and Resources Development of Shallow Lakes of Ministry of Education, College of Environment, Hohai University, Nanjing 210098, China. E-mail address: [email protected] (G. Lu).

https://doi.org/10.1016/j.scitotenv.2018.11.017 0048-9697/© 2018 Elsevier B.V. All rights reserved.

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H. Dong et al. / Science of the Total Environment 653 (2019) 1395–1406

1. Introduction Per- and polyfluoroalkylated substances (PFASs) have attracted tremendous attention in recent years for their emerging threatens to human as well as environment (Ahrens and Bundschuh, 2014). With properties of both hydrophobicity and oleophobicity, the most common PFASs, perfluorooctanoic acid (PFOA) and perfluorooctane sulfonate (PFOS), are widely applied in fluoropolymers manufacture, surfactants, textile and leather treatment, paper products and fire-fighting foams (Ahrens, 2011). Their persistence against environmental degradation and sustained release from precursors or consumer products has resulted in worldwide existence of these compounds, of which the potential hazard has not been fully elucidated (Krafft and Riess, 2015; Z. Wang et al., 2015). Following the listing of PFOS and related substances in the Stockholm Convention on persistent organic pollutants (POPs) (UNEP, 2009), PFOA and its ammonium format (APFO) were included in the Candidate List of Substances of Very High Concern for Authorisation (ECHA, 2013, 2015). However, the large production volume (90 t in 2012) in China can still generate a great quantity of PFOA emission (Li et al., 2015). To date, PFOA has been ubiquitously detected in various environmental media, such as dust, soil, surface water, sediment, fish, and even human blood (Björklund et al., 2009; T. Wang et al., 2015). Its occurrence in Antarctic and Arctic regions and in remote alpine areas, such as the Tibet plateau, indicates global contamination (Cai et al., 2012; Kwok et al., 2013; X. Wang et al., 2014). Due to its relatively high solubility (3400 mg/L), PFOA typically presents as a dissolved phase in water, which increases the risk of diffusion (Renzi, 2013). The measured levels of PFOA in water samples have been found to range from not detected (nd) to several hundreds of ng/L (Ahrens, 2011; Lu et al., 2015). The highest PFOA concentration was found in the Moehne river near Heidberg, with a value of 3640 ng/L, which was close to that detected near fluoropolymer facilities (Skutlarek et al., 2006; P. Wang et al., 2014). Additionally, Skutlarek et al. (2006) also reported extremely high PFOA concentrations in tributaries of Moehne river influenced by a wastewater treatment plant. In contrast, the average concentrations enriched in fish were at a low status of nd~b10 ng/g wet weight (ww), and the high affinity for protein has made blood an ideal reservoir for PFOA deposition (Fang et al., 2014). Thus, even with an uncertain log Kow (far below 4.5, according to European Chemicals Agency), PFOA can still exert effects on organisms through protein binding (ECHA, 2015). PFOA acts as a peroxisome proliferator in mammals, such as rodents (Li et al., 2017; Stahl et al., 2011). Nonetheless, this effect of PFOA was inconsistent in teleosts of different species or life-stages (Liu et al., 2009; Ulhaq, 2013; Yang, 2010). And few studies have focused on the potential correlations between peroxisome proliferation and mitochondrial functions, as well as the whole energy profile. In addition, PFOA can cause endocrine disruption and oxidative stress, which may also be consequences of peroxisome proliferation (Kim et al., 2010; Wei et al., 2008a; Yang et al., 2016). These effects are thought to be highly associated with some pathological alterations and phenotypic changes and need to be comprehensively researched. The crucian carp (Carassius auratus) widely inhabits freshwater environments, such as lakes, rivers and reservoirs (Ren et al., 2016). This organism has also been demonstrated as an ideal model for revealing the ecotoxicology of various chemicals in previous studies (Liu et al., 2014; Xie et al., 2016). Based on the confirmed peroxisome proliferation effects of PFOA reported in mammals, the objectives of the present study were to 1) investigate the bioconcentration patterns of PFOA in various tissues of crucian carp, including blood, kidney, liver, gills, brain and muscle, during 7 days of flow-through exposure; 2) assay the responses at the molecular level, with respect to the expression of enzymatic biomarkers and genes associated with mitochondrial function and fatty acid metabolism; 3) assess the effects of PFOA on cellular and behavioural phenotypes by liver histological examination and

swimming behavioural tests; and 4) clarify the potential relationships among different endpoints. 2. Materials and methods 2.1. Chemicals and reagents The target compound PFOA (purity N 96%) was purchased from Aladdin Industrial Corporation (Shanghai, China). Solvent-free stock solutions were prepared by dissolving different amounts of PFOA into water followed by ultrasonic dispersion. Other details of chemicals and reagents are listed in the Supplementary Information (SI). 2.2. Fish rearing and exposure Juvenile crucian carp of approximately a half-year old (fork length 12.5 ± 1.1 cm, body weight 15.1 ± 2.3 g) were obtained from Nanjing Institute of Fishery Science (Nanjing, China). To avoid the bioconcentration discrepancies between sexes, as well as differences in hormone levels, male fish were discerned by pectoral fins and selected for exposure (Hagenaars et al., 2013; Lee and Shultz, 2010). The fish were acclimated in thoroughly aerated municipal water for at least two weeks, in which the fish were fed commercial granular food (2% of mean body weight) once per day. Faeces and uneaten food were removed daily by suction. There was no mortality during the acclimation period. The ambient temperature was maintained at 24 ± 2 °C, and a natural photoperiod cycle (light/dark 12:12) was adopted. The water quality during exposure was dissolved oxygen at 7.8 ± 2.5 mg/L, a pH of 7.5 ± 0.8, and a total harness of 105.2 ± 6.5 mg CaCO3/L. Animal care and test procedures were maintained in accordance with the Guide from National Institutes of Health for the Care and Use of Laboratory Animals (NIH Publications No. 8023, revised 1978). The present study used a flow-through system with aquariums (volume 35 L) made of polypropylene (PP) materials. Ten fish were randomly allocated to each tank. Four PFOA gradient concentrations, namely 0.2, 10, 500 and 25,000 μg/L were selected to process the exposure, in which 0.2 μg/L was considered environmentally relevant, and 25,000 μg/L was set as half of the estimated no observed effect concentration (NOEC) to zebrafish embryos at 120 h post fertilization (Ahrens, 2011; Hagenaars et al., 2011). A total of 100-L solution in each tank was renewed daily. The set PFOA concentrations were fulfilled by simultaneously adjusting the fluxes of pumped water and stock solutions. The control group received dechlorinated municipal water without PFOA input. Each group had three replicate tanks. Before the exposure, six fish were taken from the acclimation aquariums and the liver mRNA were preserved and extracted for reference in gene expression quantification. On the 1st, 4th and 7th days of exposure, two random fish from each individual tank were sampled; thus, six fish were sacrificed per group at each time point. Among the six fish, three were used for determining the tissue concentrations, and the other three were used for biochemical assay (enzymatic biomarkers and transcriptional expression). The remaining twelve fish in each treatment were pooled to conduct behavioural tests. The PFOA concentration in exposure solutions was determined at every sampling time. The collected fish were first anaesthetized with MS222 (100 mg/L), and then the blood was drawn from the caudal vein. The blood samples were transferred to tubes with sodium citrate and centrifuged at 500 ×g for 20 min. The obtained supernatants (serum) were store at −20 °C. Other tissues, including liver, kidney, gills, brain and muscle, were sampled after sacrificing the fish by decapitation. Following rinsing with KCl (0.15 mol/L) and blotting, these tissues were immediately transferred to −80 °C until further use. A portion of the liver (0.1 g) was stored in paraformaldehyde (PFA, 4%) for histological observations.


2.3. Concentration measurements Water samples were extracted by using Waters Oasis WAX cartridges (Waters Co., Milford, MA, USA). The tissue samples were pretreated by an ion-pairing liquid extraction method (Pan et al. (2014). PFOA detection was performed by ultra-performance liquid chromatograph (Waters Acquity UPLC, Milford, MA, USA) in tandem with triplequadrupole mass spectrometer. Negative mode of electrospray ionization source (ESI−) was chosen. Detailed information on the sample pretreatment methods and instrumental detection conditions can be referred in the SI. Each group of twelve samples contained one procedure blank, in which no PFOA was detected. The limits of detection (LODs), calculated as three times of the signal-to-noise (S/N) ratio, were 0.041 ng/mL, 0.051 ng/mL and 0.413 ng/g for water, serum and tissue samples, respectively. The spiked PFOA (10 ng) recovery ranged from 81.3%– 99.7% for tissues (including blood serum). All PFOA concentrations in tissues were adjusted with an internal standard curve. 2.4. Biochemical determination Sampled liver or brain was mixed with nine volumes of phosphate buffer solution (pH 7.4) and mechanically homogenized. The supernatant was obtained after centrifugation (4000 ×g) for 10 min at 4 °C. Brain acetylcholinesterase (AChE) was assayed with Diagnostic Reagent Kits based on the proportionate relationship between substrate acetylcholine (ACh) depletion and chromogenic product generation. Liver carboxylesterase (CbE) was measured by a kinetic assay expressed as the formation rate of naphthyl ester in catalytic system. The enzymes were normalized by the protein contents, which were determined by Bradford Protein Assay Kits using a standard of bovine serum album (Bradford, 1976). The final units were denoted to U/mg protein, where U stood for per unit of hydrolysed ACh and increased O.D. for AChE and CbE, respectively. The acyl-CoA oxidase (ACO) activity was measured by double antibody sandwich enzyme-linked immunosorbent assay Kits (ELISA, Senbeijia BioTech Co., Nanjing, China). The process was similar to that in our previous research (Dong et al., 2018). The unit of ACO activity was defined as U/L, where U meant the generation rate of H2O2 per minute in the catalytic system. The quantification of H2O2 could be determined by a standard curve method of fluorescence intensity. Further details on the enzymatic determination are provided in the SI. A portion of the liver was used to make homogenate with boiling deionized water (1:9, m/v). Following boiling (10 min), extraction (1 min) and centrifugation (2000 ×g, 10 min), the ATP content in the supernatant was measured based on the reaction between ATP and creatine catalysed by creatine kinase. The product phosphocreatine was detected by phosphomolybdic acid colourimetry at 636 nm. The unit for ATP was expressed as μmol/g protein. 2.5. Cellular energy allocation examination The cellular energy allocation (CEA) test, which could reflect the systematic metabolic status of energy in tissues or organisms, comprised the energy available (Ea) and the energy consumption (Ec) (De Coen and Janssen, 1997). The liver tissue was first mixed and homogenized in a buffer solution, and the resulting homogenate was then separated into three parts to determine the Ea and Ec. The specific processing methods were based on those of previous studies (Nahrgang et al., 2010; Rueda-Jasso et al., 2004). Ea was calculated as the sum of carbohydrate, lipid and protein contents. Simply, the lipid concentration was spectrophotometrically determined at 340 nm against a glyceryl tripalmitate standard curve. Another portion of the homogenate was used for measuring protein and carbohydrate contents by using bovine serum albumin and glycogen, respectively, as standards. The Ec was estimated by the activity of electron transport system (ETS) at a

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mitochondrial level, which was according to the energetic reaction with NADH/NADPH as ETS stimulators and iodonitrotetrazolium as electron acceptor. More information on the CEA measurements is provided in the SI. 2.6. Gene expression assay Total RNA was extracted from the fish liver samples by using TRIzol reagent (Invitrogen Corp., Carlsbad, CA, USA) based on the manufacturer's instructions. Then it was reverse transcribed to cDNA for gene amplification. Specific primers of thirteen genes related with mitochondrial functions, lipid metabolism and apoptosis were designed by Primer Premier 5.0 software (PREMIER Biosoft, CA, USA) (Table S2). Expression measurements of these genes were accomplished by a quantitative real-time polymerase chain reaction (qRT-PCR) system. β-actin acted as an internal control. The amplification efficiencies for the target genes were in the range of 95%–100%. The relative bias of amplification efficiencies between reference gene and targeted genes were b5% (△E b 0.1) through comparing the slopes of corresponding calibration curves. The relative mRNA expression was quantified by the 2-ΔΔCt method (Livak and Schmittgen, 2001). Elaborated procedures can be found in the SI. 2.7. Histological observation The liver fixed in PFA was sequentially dehydrated with ethanol, xylene and paraffin processing. Then, the liver sample was embedded in paraffin wax and sectioned at 4 μm. After drying at 60 °C, the paraffin sections were stained with hematoxylin and eosin. The histological section was finally assessed on a Nikon Eclipse E100 microscope (Nikon instruments, Japan) equipped with a Nikon DS-U3 imaging system. 2.8. Locomotor measurements The method of swimming behavioural assay can be found in our previous research (Dong et al., 2018). The swimming activity was represented as total number of crossing the dividing lines in 600 s. 2.9. Statistical analyses Statistical analyses were performed with SPSS©IBM Statistics software (version 22.0, Chicago, IL, USA). The normality and homogeneity of the data were examined through Shapiro-Wilk's test and Levene's test, respectively. In view of the small sample size, the data was resampled with bootstrapping calculations. After that, one-way ANOVA followed by Bonferroni post hoc test was conducted for comparing the differences between treatments at different time points. The relationships between various endpoints were elucidated with Pearson's correlation coefficients. To minimize the gap with steady state and maintain uniformity with transcriptional data, these parameters were compared with ten values (i.e., on the 4th and 7th days), except for swimming activity with five values. A criterion of p b 0.05 was regarded as significant. 3. Results 3.1. Internal concentrations in tissues The PFOA concentrations in various tissues, including blood, kidney, liver, gills, brain and muscle, measured during exposure are displayed in Fig. 1. There was no PFOA detected in any tissues of the control fish. For the 0.2 μg/L treatment, PFOA was detected only in the blood, with an increasing trend, and in the liver on day 7. An increase in PFOA levels with increasing exposure concentrations was detected in all tissues, and the maximum values were typically observed on the last day. The PFOA concentrations were highest in the blood, reaching as much as

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(BCFs) were calculated for the investigated tissues; the logarithmic values are presented in Fig. 2. In general, the BCF values decreased with increasing exposure concentrations. The BCFs at two time points (days 4 and 7) ranged from 1.1–60.4 for blood, 0.7–29.1 for liver, 0.2–5.9 for brain, and 0.1–6.5 for muscle. The BCF ranking of intertissues was similar to the order of internal concentrations. 3.2. Biochemical responses

Fig. 1. PFOA concentrations in tissues after 1, 4 and 7 days of exposure with different treatments (Bl-blood, K-kidney, L-liver, G-gill, Br-brain, M-muscle). Blank spots means not detected in corresponding tissues. Values are presented as mean ± standard deviation (SD) (n = 3).

23,688 ng/mL (25,000 μg/L treatment on day 7). In contrast, PFOA enrichment in muscle tissue was much lower, with a terminal internal concentration of 2935.5 ng/g under 25,000 μg/L PFOA. The global concentrating order for tissues was blood N kidney ≥ liver N gill N brain N muscle. By combining the actual exposure concentrations (Table S3 in the SI) with the tissue concentrations of PFOA, the bioconcentration factors

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significant (p N 0.05). The main constituent, viz. lipids (accounting for 76.3%–94.0%), contributed to the major reduction in the exposure groups (data not shown). Most of the CEA values in PFOA treatments were lower than those in control, excepted for those determined on

the first day. However, the inner-group comparisons at individual time point did not show any significant differences. Notably, a decrease in the CEA value with increasing PFOA concentrations (≥10 μg/L) was observed on days 4 and 7.

Table 1 Energy available (Ea) and cellular energy allocation (CEA) in fish livers at different time. Data are presented as mean ± standard deviation (SD) (n = 3). ⁎ indicates values were significant different from that in control (p b 0.05). Treatment

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(μg/L)

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25.06 ± 4.62 19.67 ± 1.23 18.08 ± 3.50 19.14 ± 2.46 16.69 ± 3.39

26.96 ± 5.25 19.45 ± 4.48 18.90 ± 1.33 17.62 ± 2.06 17.02 ± 4.14

26.89 ± 3.36 22.67 ± 2.25 18.13 ± 3.30⁎ 19.65 ± 2.89 19.51 ± 3.46

2.64 ± 0.16 3.11 ± 0.19 2.65 ± 0.51 3.16 ± 0.41 3.79 ± 0.77

2.71 ± 0.53 3.00 ± 0.69 2.71 ± 0.19 2.57 ± 0.30 2.07 ± 0.50

2.70 ± 0.34 2.79 ± 0.25 2.59 ± 0.43 2.57 ± 0.38 2.28 ± 0.40


3.3. Gene expression The gene expression of respiratory chain complexes I\\V in the fish liver (mitochondria) on days 4 and 7 of exposure is graphed in Fig. 4. Complex I - NADH dehydrogenase subunit 1 (nd-1) was limitedly upregulated and increased to overtly higher levels (compared with that in the control) at 10 μg/L PFOA on day 4 and at 0.2 μg/L on day 7. Complex II - succinate dehydrogenase (sdh) primarily exhibited transcriptional repression, with a maximum downregulation of 60% at 0.2 μg/L treatment on day 4. After treatment at 25000 μg/L, sdh expression was restored to a level close to that in the control. Complex III - cytochrome b (cytb) was significantly upregulated at 10 μg/L on day 4, but on day 7, cytb showed a decreasing trend with increasing PFOA concentrations. The mRNA expression levels of complex IV - cytochrome c oxidase subunit 1 (cox-1) were markedly downregulated with PFOA treatments in all groups on day 4 and in the low dose groups (0.2 and 10 μg/L) on day 7 (69%–81% of the control). Complex V - ATP synthase F0 subunit 6 (atp06) showed dose-dependent decline in expression on both days 4 and 7, with maximum downregulation of 40% (p N 0.05) and 50% (p b 0.05), respectively, compared with the corresponding controls. Gene expressions of two selected mitochondrial permeability transition pore (MPTP) proteins were also examined and illustrated (Fig. 4). Different dose-dependent patterns for adenine nucleotide translocator (ant) were detected on day 4 (an increase with increasing PFOA concentrations) and day 7 (bell-shape). The summits were obtained at 25000 μg/L on day 4 (1.8-fold, p b 0.05) and 10 μg/L on day 7 3.2

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(1.5-fold, p b 0.05). Voltage-dependent anion channel 1 (vdac) was similarly upregulated on both days, with an initial increase and subsequent decrease. The maximum upregulation reached 5.2-fold at 10 μg/L on day 4 and 3.7-fold at 500 μg/L PFOA on day 7. Compared with the control, the mRNA expression levels of peroxisome proliferator activated receptor alpha (ppar-α) gradually increased with increasing PFOA concentrations, showing significantly higher levels when exposed to doses ≥500 μg/L. Notably, on day 7, ppar-α expression was attenuated after 25,000 μg/L PFOA treatment (Fig. 5). For carnitine palmitoyl transferase 1A (cpt-1), the expression remained almost the same, with some mild upregulation (p N 0.05) after treatments ≥10 μg/L. The mRNA expression of fatty acid-binding protein, liver-like (l-fabp) was also increased, with maximum induction after treatment at 500 μg/L. The uncoupling protein 2 (ucp-2) expression was significantly upregulated at PFOA concentrations ≥10 μg/L, except for the 25,000 μg/L treatment on day 4. Additionally, the transcriptional levels of two apoptosis-linked genes bcl-2 and bax showed contrasting changes (see Fig. S1 in the SI). The bcl-2 expression was mostly downregulated, while bax expression was dose-dependently upregulated, with maximums levels observed at 500 μg/L. 3.4. Liver histology Based on histological observation, the pathological alterations at cellular level in response to PFOA are illustrated in Fig. 6. In the livers of fish under 0.2 μg/L PFOA treatment, there were no obvious morphological

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Exposure concentrations (µg/L) Fig. 4. The mRNA expression levels of respiratory chain complex I\ \V and related MPTP proteins in male Carassius auratus of control and PFOA treatments on day 4 and 7. Data are plotted as mean ± standard deviation (SD) (n = 3). Significant differences are denoted by different letters for each indicator (p b 0.05).


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Fig. 5. The mRNA expression levels of various indexes concerned with energetic metabolism in male Carassius auratus of control and PFOA treatments on day 4 and 7. A: ppar-α; B: cpt-1; C: l-fabp; D: ucp-2. Data are plotted as mean ± standard deviation (SD) (n = 3). Significant differences are denoted by different letters (p b 0.05).

Fig. 6. Liver histopathology of male Carassius auratus after 7 days of PFOA exposure. Liver sections were stained with H&E under a magnification of 400×, and scale bar is equal to 30 μm. A: the control; B: 10 μg/L; C: 500 μg/L; D: 25000 μg/L. Arrow: nuclei pycnosis; asterisk: vacuolar degradation; triangle: nuclei decomposition; and rhombus: eosinophilic hyaline droplets.

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changes compared with the control. However, in the 10 μg/L treatment group, some anomalies of hepatocytes, e.g., nuclei enlargement or pyknosis and vacuolar degradation, were observed. Exposure to 500 μg/L PFOA elicited severe nuclei enlargement and vacuolar degradation, accompanied by the appearance of eosinophilic hyaline droplets. At 25000 μg/L PFOA, the frame structure of some hepatocytes was nearly decomposed. 3.5. Swimming performance The grid movements of the fish showed a decreasing trend with an increasing PFOA concentration gradient (Fig. 7). Swimming activity was significantly suppressed (46.3%–56.8%) at the highest concentration of PFOA (p b 0.05). 4. Discussion 4.1. Bioconcentration patterns of PFOA As it is increasingly detected and identified as the primary PFAS in natural water, PFOA can readily come in contact and subsequently exert its effects on aquatic organisms. In the present study, PFOA was shown to preferentially concentrate in blood, kidney and liver, demonstrating a certain distinction in the gills, brain and muscle. This finding is consistent with the observations of previous exposure studies, in which fish blood and liver acted as the main sinks for PFOA (Giari et al., 2016; Martin et al., 2003a; Rotondo et al., 2018). These results were also evidenced by the detection of PFOA in blood and liver tissues under environmentally relevant PFOA concentration (0.2 μg/L). The PFOA enrichment in the blood is largely attributed to its proteinophilic feature, as plasma has the highest protein content (mainly albumin) among these tissues (Liu et al., 2011). The highly vascularized and diffused properties of the liver, as well as the abundant fatty acid binding proteins (FABPs) may explain the relative high PFOA concentrations appeared (Giari et al., 2016; Ng and Hungerbühler, 2013; Ulhaq et al., 2015). Moreover, enterohepatic recirculation might be among other pivotal reasons to account for this effect (Martin et al., 2003b). The comparable bioconcentration in kidney and liver tissues was consistent with the findings of Mortensen et al. (2011), reflecting the high perfusion of blood (Martin et al., 2003a). However, the mechanism of renal excretion

Fig. 7. Swimming activity of male Carassius auratus after 7 days of PFOA exposure. The range of each box is from 25th (down) to 75th (up) percentile. The whiskers below and above the boxes refer to the minimum and maximum values respectively. The horizontal lines and squares inside the boxes display the median and mean values in each individual treatment. Significant differences are represented by different letters (p b 0.05).

and reuptake of PFOA is worthy of further clarification (Armitage et al., 2017). Although the gills directly contact PFOA in exposure solutions, permeation was the rate-limiting factor, resulting in lower PFOA concentrations in these tissues (Lee and Shultz, 2010). The elimination of the gills cannot be ignored, as pointed out by Martin et al. (2003b). The PFOA profile in the brain was similar to that in the gills, which demonstrated that PFOA could overcome the blood-brain barrier and likely induce the subsequent neurotoxicity (Ulhaq et al., 2015). Muscle was the least PFOA-enriched tissue, indicating a relatively low affinity, and this phenomenon may originate from the dynamic distribution order in tissues (Ng and Hungerbühler, 2013). A similar distribution profile in fish tissues was also reported by previous field investigations (Fang et al., 2014; Giari et al., 2015). Under environmentally relevant PFOA exposure, the blood and liver concentrations after 7 days (25.9 ng/mL and 5.1 ng/g, respectively) were comparable with those in wild fish samples from Taihu Lake (Fang et al., 2014), but were much higher than those in European chub (Leuciscus cephalus) (Labadie and Chevreuil, 2011). At PFOA treatments ≥10 μg/L, the concentrations in different tissues were ca. 0.2–3.1 orders of magnitude greater than those in captured teleosts (Lam et al., 2014; Pan et al., 2014). The tissue-specific BCFs were changed at an overt range. Generally, the blood had higher BCFs than other tissues, suggesting a stronger affinity to PFOA. Higher BCF values were typically obtained in low PFOA treatments, consistent with the observations of the steady-state bioconcentration of common carp (Cyprinus carpio) (Inoue et al., 2012). The BCFs of the tissues examined here were also numerically similar to those in frog tadpoles (Hoover et al., 2017), as well as in rainbow trout (Martin et al., 2003a) from PFOA laboratory aqueous exposure studies. 4.2. Molecular responses Molecular level responses, including enzymatic responses and gene expression, are crucial indices to characterize the mode of action (MoA) of chemicals. With respect to neural effects, ACh is an important neurotransmitter included in the cholinergic system and closely associated with the neurodevelopment of aquatic vertebrates. AChE hydrolyses ACh to choline and acetate in the synaptic cleft (Lushchak, 2011). Thereby, AChE inactivation leads to the over-accumulation of ACh, which is a typical pathway of xenobiotics to exert neurotoxicity (Richendrfer et al., 2014). Consistent with this, the decreased AChE activity observed in the present study probably demonstrated PFOA acted as an AChE inhibitor in fish, as some organophosphate compounds, and pharmaceuticals did (Liu et al., 2014; Richendrfer et al., 2014). However, another flow-through exposure study has shown that the AChE activity in the common carp was not influenced by PFOA (Kim et al., 2010). Jeong et al. (2016) further reported a stimulatory effect of PFOS on AChE of Daphnia magna offsprings. The difference of sensitivity in different species and toxicity of PFASs with different Cchain and functional groups might account for the varied results. Thus, the inhibitory effect and specific MoA on piscine AChE brought by PFOA are still inconclusive and require further investigation. Similar to AChE, CbE is a B-type serine esterase. CbE participates in the phase I metabolism of xenobiotics, as well as the hydrolysis of endogenous esters and thioesters associated with lipid metabolism (Lian et al., 2018). In the present study, CbE was generally and markedly stimulated by PFOA, which could be interpreted as a detoxifying mechanism against toxicant-related stresses that would in turn confer protection to cholinesterases (e.g., AChE) (Barron et al., 1999). Moreover, peroxisome proliferators induce CbE in animals, indicating the potential regulation of the peroxisome proliferator response element (PPRE) on CbE activity (Derbel et al., 1996; Moody et al., 1992). PFOA has previously been implicated as a peroxisome proliferator (Wei et al., 2008b; Yang, 2010), which can cause increase of CbE through increasing the liver smooth endoplasmic reticulum (Moody et al., 1992). Consistently, the hypolipidemic drug gemfibrozil, also identified as a peroxisome


proliferator, enhanced microsomal CbE activity in Solea senegalensis (Sole et al., 2014). While the decline of CbE with time might be originated from an adaptive response in fish. The peroxisome proliferation effect can be quantified by ACO, which acts as a rate-limiting enzyme in fatty acid peroxisomal β-oxidation (Liu et al., 2008). The dose-dependent induction of ACO activity caused by PFOA demonstrated the disruption of lipid metabolism in PFOA toxic pathways. Similarly, Oakes et al. (2005) showed that PFOS consistently increased fatty ACO in different fish species. In contrast, another study with 39 days of exposure showed that fatty ACO activity was elevated in fathead minnow (Pimephales promelas) treated with 300 μg/L of PFOA but attenuated at higher concentrations (Oakes et al., 2004). Stahl et al. (2011) attributed this to the adaptive downregulation of hepatic peroxisome proliferation under high cumulative concentrations. Accordingly, the phenomenon in the present study might reflect some hysteresis of ACO activity compared with transcriptional response (see expression of ppar-α below). For eukaryotic organisms, mitochondria are the main sites for ATP generation through oxidative phosphorylation (Yeh, 2017). In the manner of respiratory control, this process is highly associated with ETS activity, which provides a direct assessment of Ec (De Troch et al., 2013; Smith et al., 2012). The decreased ETS activity observed here might be caused by mitochondrial dysfunction under PFOA treatments. Many chemicals including PFASs induce multiple mitochondrial defects, such as oxidative stress, disturbance of membrane integrity, and interference with specific respiratory complexes of the ETS (van Boxtel et al., 2008; Yeh, 2017). The observation showed some similarity with the ETS inhibition by PFOA on zebrafish (Hagenaars et al., 2013), except for the certain recovery as time proceeding and that in 25,000 μg/L PFOA treatment. It was possibly derived from the susceptibility of ETS and a self-restoration process as well (Oh et al., 2013). These studies also highlighted ATP depletion in vivo, suggesting some causal mechanisms based on the fact that ATP synthesis links directly with ETS efficiency. Therefore, the gradually decreasing ATP contents influenced by PFOA in the present study could in turn reflect the undermined ETS activity to some extent. Additionally, PFOA suppressed the Ea, particularly the lipids, which could be vital substrates for ATP generation. This effect was likely associated with a direct reduction of fatty acid biosynthesis (Wei et al., 2008b). Other processes, such as uncoupling respiration, might also spend the energy reserves without increasing ATP production. Despite the inhibition of Ec (ETS activity), the calculated CEA exhibited a decreasing trend after 4 days in fish treated at higher PFOA concentrations (≥10 μg/L), indicating a reduced energy budget. This finding might represent a decreasing overall condition of the fish according to Rueda-Jasso et al. (2004). With respect to the components of the ETS, the transcriptional levels of respiratory chain complexes I\\V were examined to specify the PFOA effects on the related oxidative phosphorylation. ETS is initiated with electrons from NADH- and FADH2-linked substrates via NADH dehydrogenase (complex I) and succinate dehydrogenase (complex II), respectively. Under a low energy status (refer to the relatively low Ea value described above), the selectively enhanced nd-1 mRNA expression likely resulted from a compensatory mechanism to strengthen electron donation (Liao et al., 2016). This effect was similar to the proteomic alterations in PFOA-treated zebrafish, but different from gene transcriptional aspect (Hagenaars et al., 2013). In contrast, the transcriptional levels of sdh were downregulated in most cases, consistent with the observations of Yang et al. (2016). Yeh (2017) summarized that the activity of complex II was well correlated with the oxygen consumption rate (OCR), which represented the status of ETS. Thus, the effects observed in the present study might be a direct contributor to the declined ATP generation, as confirmed by the downregulation of complex V. Even though the erratic disturbance on cytochrome b (complex III) further magnified the unpredictability of electron transport efficiency, we could still identified a decreasing trend of cytb expression on day 7. Furthermore, the globally reduced mRNA expression levels of cox-1 could reflect a certain

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interference with the mitochondrial function through PFOA. According to Berthiaume and Wallace (2002), the activity of complex IV was likely a proper indicator of mitochondrial number. Therefore, the inhibited expression observed in the present study might represent a discounted mitochondria turnover. The expression of sdh and cox-1 were both characterized with a paralleled recovery in high PFOA treatments, which demonstrated an adaptive mechanism as that on ETS activity. Overall, the alterations suggested attenuated ETS efficiency and depressed mitochondrial function. The mitochondrial membrane permeability is tightly associated with the efficiency of energy generation and a cascade of ecological functions. Induced opening of the MPTP can lead to increased membrane permeability and reduced transmembrane potential (ΔΨm), which facilitate the uncoupling of oxidative phosphorylation and contribute to undermine ATP synthesis (Liu et al., 2008). Consistent with a previous study (Hagenaars et al., 2013), the expression levels of both MPTP proteins were upregulated, indicating an obvious enhancement of membrane fluidity. Moreover, the induced opening of the MPTP accelerates the release of calcium, cytochrome c and apoptotic proteins, further resulting in cell apoptosis (Hu et al., 2016). PFOA exerts its hepatotoxicity in rodents mainly via peroxisome proliferation (Li et al., 2017). By ligand binding, the principal mediator peroxisome proliferator activated receptor α (PPARα) plays a crucial role in fatty acid metabolism (Stahl et al., 2011). The upregulation of the corresponding gene ppar-α following PFOA treatment well demonstrated PFOA as a PPARα agonist in the crucian carp. This finding could be supported by experiments in different teleost species (Fang et al., 2010; Yang, 2010). While Ulhaq (2013) did not observe significant changes for ppar-α mRNA expression in zebrafish but with a decreasing trend. Inhibitory effects were also reported in fish exposed to other PFASs, such as perfluorononanoic acid and perfluorododecanoic acid (Liu et al., 2008; Zhang et al., 2012a). The differences in species, maturity of the fish and carbon length of the PFASs would account for the inconsistent results. Notably, the small decline of ppar-α expression at 25000 μg/L PFOA on day 7 corresponded to an adaptive response as mentioned previously. The ppar-α target gene, cpt-1 mediates the catalysed conversion of acyl-CoA into acyl-carnitine for entry into the mitochondrial matrix (Liu et al., 2008). The transcriptional levels of cpt-1 were moderately elevated but not statistically different from those in the control. This finding is similar to those reported in previous studies, in which the mRNA expression of cpt-1 was well correlated with ppar-α expression (Jin et al., 2011; Zhang et al., 2012a). This result could reflect an inclination of fatty acid mimics into mitochondria. As a structural analogue of endogenous fatty acids, PFOA can competitively bind to liver FABPs, which play pivotal roles in the uptake, sequestering and transport of fatty acids, and these proteins also interact with other transport and enzyme systems (Liu et al., 2015). Once transported to nucleus, PFOA activates PPARα, which forms heterodimers with retinoid X receptors and then binds to PPREs, such as l-fabp gene to in turn stimulate its transcription (Zhang et al., 2012b). Thus, the elevated expression of l-fabp mRNA might reflect the increasing entry of PFOA into the cytoplasma, and the activation of ppar-α and the interference of fatty acids binding are potential mechanisms for PFOA-induced peroxisome proliferation (Wei et al., 2008a). In addition, higher concentrations of PFOA (≥10 μg/L) markedly increased the mRNA expression levels of ucp-2, another target gene of ppar-α. As mitochondrial anion carrier proteins, uncoupling proteins (UCPs) promote proton leak across the inner membrane and hence diminish the membrane potential, further curtailing the production of reactive oxygen species (ROS) (Liu et al., 2015; Liu et al., 2008). The observations in the present study might be attributed to both ppar-α upregulation and a defensive strategy to reduce the enhanced ROS (Liu et al., 2008). Moreover, the increased uncoupling of oxidative phosphorylation would lead to a decrease in ATP synthesis, similar to the results of the present study. The expression of bcl-2 can also represent an antioxidant process to prevent apoptosis, while bax has the opposite function of pro-apoptosis

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(Yang et al., 2014). Together, the bax/bcl-2 ratio is regarded as a vital determinant of susceptibility to apoptosis (Yang et al., 2016). Thus, the increased bax but decreased bcl-2 expression in the present study might represent an ongoing apoptotic process. 4.3. Effects on cellular and behavioural phenotypes In the present study, PFOA induced cellular lesions on hepatocytes, which became more severe with increasing exposure concentrations, and nuclei enlargement and pyknosis were observed with intensifying vacuolation, showing parallel phenomena with different PFASs exposure in rats and fish (Fang et al., 2012; Giari et al., 2016; Liu et al., 2008; Wei et al., 2008a). Giari et al. (2015) showed that the likely reversible vacuolation might be a consequence of the alteration of membrane permeability. The eosinophilic hyaline droplets resulted from the expression of pro-inflammation cytokines. Moreover, the karyolysis and decomposition of the cellular structure in high PFOA treatments may indicate the activation of apoptosis. Swimming activity has been demonstrated as an useful tool to assess the toxic effects on aquatic organisms caused by xenobiotics (Xia et al., 2014; Xie et al., 2016). Correlated with individual fitness, swimming performance can directly impact activities, such as social interaction, attack and capture of prey, and predator evasion (Pessoa et al., 2011). The gradually declining swimming activity observed in the present study paralleled the results in previous fish experiments with various chemicals (Liu et al., 2016; Macaulay et al., 2015; Xia et al., 2014). This effect may result from multiple pathways, for instance, AChE inhibition and the resulting neurotoxicity, reduced energy synthesis efficiency and consumption, muscular damage induced by oxidative stress (Abe et al., 2018; Macaulay et al., 2015). 4.4. Relationships among endpoints Pearson's correlation analysis was conducted to determine potential correlations between tissue PFOA concentrations, brain AChE activity, liver enzymatic and transcriptional responses, and behavioural quantified parameter (see Table S4 in the SI). A significantly negative correlation between PFOA concentrations in brain and AChE activity (r = −0.979, p b 0.01), with close correlations between internal liver concentrations and enzymatic responses, reflected a direct effect on the biochemical signature of the crucian carp brought by PFOA bioconcentration. Expectedly, the PFOA enriched in blood serum was also well correlated with these enzymes, as abundant proteins in the blood mediate FPOA binding and entry into target tissues, such as the liver (Ng and Hungerbühler, 2013). However, ETS activity at the protein level and its components (respiratory chain complexes I-IV) at the transcriptional level did not show strong correlations with liver PFOA concentrations, indicating that the ETS might be sensitive to various factors caused by xenobiotic stressors, such as oxygen consumption efficiency, membrane integrity, and the enhanced uncoupling process (van Boxtel et al., 2008; Yeh, 2017). In contrast, PFOA exposure interfered with ATP synthesis and subsequently caused a declined in ATP status, from obviously inverse relations between PFOA concentrations in the liver and the mRNA expression of atp06 (r = −0.957, p b 0.01) and ATP contents (r = −0.958, p b 0.01). This effect likely resulted from peroxisome proliferation, which was confirmed by ppar-α activation with increasing PFOA concentration (r = 0.851, p b 0.01) in the present study. The close negative relationships between decreased swimming activity and internal tissue concentrations should be regarded as an adverse outcome of PFOA on spontaneous behaviour. Within the enzymatic parameters, CbE activity was negatively correlated with AChE activity (r = −0.754, p b 0.05) but positively associated with ACO activity (r = 0.658, p b 0.05), illustrating that CbE induction might be driven by elevated stress or result from peroxisome proliferation (Barron et al., 1999; Li, 2008). The ppar-α expression was highly correlated with the marker enzyme ACO, as well as the synthesis and

status of ATP, suggesting that peroxisome proliferation consistently suppressed energy metabolism. The energy depletion might also be a consequence of strengthened uncoupling pathway from the markedly negative correlation between ucp-2 expression and ATP contents (r = −0.706, p b 0.05). Furthermore, positively correlated mRNA expression levels of sdh and cox-1 with ETS activity might indicate undermined mitochondrial function, as noted in previous studies (Cambier et al., 2009; Yang et al., 2016). This effect was also tightly associated with the integrity of the mitochondrial membrane, even with approximately significant correlations between the mRNA expression levels of MPTP proteins and the ETS (p = 0.053 for ant, p = 0.078 for vdac). The close correlation between cpt-1 and vdac expression might reflect more opened MPTPs, which likely result from the reduced transmembrane ΔΨm due to increased fatty acid mimics that enter mitochondria. This phenomenon could also be induced by the upregulated expression of ucp-2, for the corresponding protein facilitated the proton leak. From the significant and positive correlation between ppar-α and bax/bcl-2 (r = 0.750, p b 0.05), the activation of apoptosis has been correlated with a peroxisome proliferation effect, consistent with histological observations. The aggravated lesions were likely dosedependent because of the correlation between bax/bcl-2 and liver PFOA concentrations. There were close relationships between swimming performance and brain AChE (r = 0.934, p b 0.05) and liver ATP contents (r = −0.887, p b 0.05), indicating that the behavioural parameter was a reflection of multi-effects concerned with neural toxicity, global energy profile, etc. However, this finding only provided a holistic estimation since some studies have indicated a specific correlation between OCR and critical swimming speed (Lee, 2003). 5. Conclusions In the present study, the bioconcentration and biochemical and phenotypic effects of PFOA in male crucian carp (Carassius auratus) were investigated by a flow-through system. PFOA was preferentially concentrated in blood and some diffused tissues, such as the liver and kidney, demonstrating a proteinophilic property. The enzymatic responses suggested neurotoxicity and peroxisome proliferation, accompanied by undermined ETS activity and declined ATP turnover. The transcriptional alterations of related genes showed mitochondrial dysfunction, which might be associated with peroxisome proliferation at both protein and transcriptional levels. With respect to the phenotypes, histological pathology and reduced swimming activity reflected a direct adverse effect of PFOA. Taken together, the results of the present study confirmed the impact of PFOA on energy metabolism and revealed potential correlations between biochemical effects and phenotypes. Conflict of interest The authors declare that they have no conflict of interest. Ethical approval All applicable international, national, and/or institutional guidelines for the care and use of animals were followed. This article does not contain any studies with human participants performed by any of the authors. Informed consent Informed consent was obtained from all individual participants included in the study. Acknowledgements This study was supported by the National Natural Science Foundation of China (Grant 51879228), the National Science Funds for Creative


Research Groups of China (Grant 51421006), the Fundamental Research Funds for the Central Universities (Grant 2018B43614), the Program for Scientific Research Innovation Team in Colleges and Universities of Tibet Autonomous Region, and the Priority Academic Program Development of Jiangsu Higher Education Institutions. Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi. org/10.1016/j.scitotenv.2018.11.017. References Abe, F.R., Soares, A., Oliveira, D.P., Gravato, C., 2018. Toxicity of dyes to zebrafish at the biochemical level: cellular energy allocation and neurotoxicity. Environ. Pollut. 235, 255–262. Ahrens, L., 2011. Polyfluoroalkyl compounds in the aquatic environment: a review of their occurrence and fate. J. Environ. Monit. 13, 20–31. Ahrens, L., Bundschuh, M., 2014. Fate and effects of poly- and perfluoroalkyl substances in the aquatic environment: a review. Environ. Toxicol. Chem. 33, 1921–1929. Armitage, J.M., Erickson, R.J., Luckenbach, T., Ng, C.A., Prosser, R.S., Arnot, J.A., Schirmer, K., Nichols, J.W., 2017. Assessing the bioaccumulation potential of ionizable organic compounds: current knowledge and research priorities. Environ. Toxicol. Chem. 36, 882–897. Barron, M.G., Charron, K.A., Stott, W.T., Duvall, S.E., 1999. Tissue carboxylesterase activity of rainbow trout. Environ. Toxicol. Chem. 18, 2506–2511. Berthiaume, J., Wallace, K.B., 2002. Perfluorooctanoate, perflourooctanesulfonate, and Nethyl perfluorooctanesulfonamido ethanol; peroxisome proliferation and mitochondrial biogenesis. Toxicol. Lett. 129, 23–32. Björklund, J.A., Thuresson, K., de Wit, C.A., 2009. Perfluoroalkyl compounds (PFCs) in indoor dust: concentrations, human exposure estimates, and sources. Environ. Sci. Technol. 43, 2276–2281. Bradford, M.M., 1976. A rapid and sensitive method for the quantitation of microgram quantities of protein. Anal. Biochem. 72, 248–254. Cai, M., Yang, H., Xie, Z., Zhao, Z., Wang, F., Lu, Z., Sturm, R., Ebinghaus, R., 2012. Per- and polyfluoroalkyl substances in snow, lake, surface runoff water and coastal seawater in Fildes peninsula, King George Island, Antarctica. J. Hazard. Mater. 209-210, 335–342. Cambier, S., Benard, G., Mesmer-Dudons, N., Gonzalez, P., Rossignol, R., Brethes, D., Bourdineaud, J.P., 2009. At environmental doses, dietary methylmercury inhibits mitochondrial energy metabolism in skeletal muscles of the zebra fish (Danio rerio). Int. J. Biochem. Cell Biol. 41, 791–799. De Coen, W.M., Janssen, C.R., 1997. The use of biomarkers in Daphnia magna toxicity testing. IV. Cellular energy allocation: a new methodology to assess the energy budget of toxicant-stressed Daphnia populations. J. Aquat. Ecosyst. Stress. Recover. 6, 43–55. De Troch, M., Reubens, J.T., Heirman, E., Degraer, S., Vincx, M., 2013. Energy profiling of demersal fish: a case-study in wind farm artificial reefs. Mar. Environ. Res. 92, 224–233. Derbel, M., Hosokawa, M., Satoh, T., 1996. Differences in the induction of carboxylesterase RL4 in rat liver microsomes by various perfluorinated fatty acids, metabolically inert derivatives of fatty acids, metabolically inert derivatives of fatty acids. Biol. Pharm. Bull. 19, 765–767. Dong, H., Lu, G., Yan, Z., Liu, J., Yang, H., Nkoom, M., 2018. Bioconcentration and effects of hexabromocyclododecane exposure in crucian carp (Carassius auratus). Ecotoxicology 27, 313–324. ECHA, 2013. Inclusion of Substances of Very High Concern in The Candidate List. European Chemicals Agency, Helsinki, p. 4. ECHA, 2015. Background Document to the Opinion on the Annex XV Dossier Proposing Restrictions on Perfluorooctanoic Acid (PFOA), PFOA Salts and PFOA-related Substances (in: RAC., SEAC. (Eds.), Helsinki). Fang, X., Wei, Y., Liu, Y., Wang, J., Dai, J., 2010. The identification of apolipoprotein genes in rare minnow (Gobiocypris rarus) and their expression following perfluorooctanoic acid exposure. Comp. Biochem. Physiol. C. Toxicol. Pharmacol. 151, 152–159. Fang, X., Gao, G., Xue, H., Zhang, X., Wang, H., 2012. In vitro and in vivo studies of the toxic effects of perfluorononanoic acid on rat hepatocytes and Kupffer cells. Environ. Toxicol. Pharmacol. 34, 484–494. Fang, S., Zhao, S., Zhang, Y., Zhong, W., Zhu, L., 2014. Distribution of perfluoroalkyl substances (PFASs) with isomer analysis among the tissues of aquatic organisms in Taihu Lake, China. Environ. Pollut. 193, 224–232. Giari, L., Guerranti, C., Perra, G., Lanzoni, M., Fano, E.A., Castaldelli, G., 2015. Occurrence of perfluorooctanesulfonate and perfluorooctanoic acid and histopathology in eels from north Italian waters. Chemosphere 118, 117–123. Giari, L., Vincenzi, F., Badini, S., Guerranti, C., Dezfuli, B.S., Fano, E.A., Castaldelli, G., 2016. Common carp Cyprinus carpio responses to sub-chronic exposure to perfluorooctanoic acid. Environ. Sci. Pollut. Res. 23, 15321–15330. Hagenaars, A., Vergauwen, L., De Coen, W., Knapen, D., 2011. Structure-activity relationship assessment of four perfluorinated chemicals using a prolonged zebrafish early life stage test. Chemosphere 82, 764–772. Hagenaars, A., Vergauwen, L., Benoot, D., Laukens, K., Knapen, D., 2013. Mechanistic toxicity study of perfluorooctanoic acid in zebrafish suggests mitochondrial dysfunction to play a key role in PFOA toxicity. Chemosphere 91, 844–856. Hoover, G.M., Chislock, M.F., Tornabene, B.J., Guffey, S.C., Choi, Y.J., De Perre, C., Hoverman, J.T., Lee, L.S., Sepúlveda, M.S., 2017. Uptake and depuration of four per/polyfluoroalkyl

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substances (PFASS) in northern leopard frog Rana pipiens tadpoles. Environ. Sci. Technol. Lett. 4, 399–403. Hu, Y., Chen, J., Fan, H., Xie, P., He, J., 2016. A review of neurotoxicity of microcystins. Environ. Sci. Pollut. Res. 23, 7211–7219. Inoue, Y., Hashizume, N., Yakata, N., Murakami, H., Suzuki, Y., Kikushima, E., Otsuka, M., 2012. Unique physicochemical properties of perfluorinated compounds and their bioconcentration in common carp Cyprinus carpio L. Arch. Environ. Contam. Toxicol. 62, 672–680. Jeong, T.Y., Yuk, M.S., Jeon, J., Kim, S.D., 2016. Multigenerational effect of perfluorooctane sulfonate (PFOS) on the individual fitness and population growth of Daphnia magna. Sci. Total Environ. 569–570, 1553–1560. Jin, Y., Zheng, S., Pu, Y., Shu, L., Sun, L., Liu, W., Fu, Z., 2011. Cypermethrin has the potential to induce hepatic oxidative stress, DNA damage and apoptosis in adult zebrafish (Danio rerio). Chemosphere 82, 398–404. Kim, W.K., Lee, S.K., Jung, J., 2010. Integrated assessment of biomarker responses in common carp (Cyprinus carpio) exposed to perfluorinated organic compounds. J. Hazard. Mater. 180, 395–400. Krafft, M.P., Riess, J.G., 2015. Per- and polyfluorinated substances (PFASs): environmental challenges. Curr. Opin. Colloid Interface Sci. 20, 192–212. Kwok, K.Y., Yamazaki, E., Yamashita, N., Taniyasu, S., Murphy, M.B., Horii, Y., Petrick, G., Kallerborn, R., Kannan, K., Murano, K., Lam, P.K., 2013. Transport of perfluoroalkyl substances (PFAS) from an arctic glacier to downstream locations: implications for sources. Sci. Total Environ. 447, 46–55. Labadie, P., Chevreuil, M., 2011. Partitioning behaviour of perfluorinated alkyl contaminants between water, sediment and fish in the Orge River (nearby Paris, France). Environ. Pollut. 159, 391–397. Lam, N.-H., Cho, C.-R., Lee, J.-S., Soh, H.-Y., Lee, B.-C., Lee, J.-A., Tatarozako, N., Sasaki, K., Saito, N., Iwabuchi, K., Kannan, K., Cho, H.-S., 2014. Perfluorinated alkyl substances in water, sediment, plankton and fish from Korean rivers and lakes: a nationwide survey. Sci. Total Environ. 491–492, 154–162. Lee, C.G., 2003. The effect of temperature on swimming performance and oxygen consumption in adult sockeye (Oncorhynchus nerka) and coho (O. kisutch) salmon stocks. J. Exp. Biol. 206, 3239–3251. Lee, J.J., Shultz, I.R., 2010. Sex differences in the uptake and disposition of perfluorooctanoic acid in fathead minnows after oral dosing. Environ. Sci. Technol. 44, 491–496. Li, M.H., 2008. Effects of nonylphenol on cholinesterase and carboxylesterase activities in male guppies (Poecilia reticulata). Ecotoxicol. Environ. Saf. 71, 781–786. Li, L., Zhai, Z., Liu, J., Hu, J., 2015. Estimating industrial and domestic environmental releases of perfluorooctanoic acid and its salts in China from 2004 to 2012. Chemosphere 129, 100–109. Li, K., Gao, P., Xiang, P., Zhang, X., Cui, X., Ma, L.Q., 2017. Molecular mechanisms of PFOAinduced toxicity in animals and humans: implications for health risks. Environ. Int. 99, 43–54. Lian, J., Nelson, R., Lehner, R., 2018. Carboxylesterases in lipid metabolism: from mouse to human. Protein & Cell 9, 178–195. Liao, K., Yan, J., Mai, K., Ai, Q., 2016. Dietary lipid concentration affects liver mitochondrial DNA copy number, gene expression and DNA methylation in large yellow croaker (Larimichthys crocea). Comp. Biochem. Physiol. B Biochem. Mol. Biol. 193, 25–32. Liu, Y., Wang, J., Wei, Y., Zhang, H., Xu, M., Dai, J., 2008. Induction of time-dependent oxidative stress and related transcriptional effects of perfluorododecanoic acid in zebrafish liver. Aquat. Toxicol. 89, 242–250. Liu, Y., Wang, J., Liu, Y., Zhang, H., Xu, M., Dai, J., 2009. Expression of a novel cytochrome P450 4T gene in rare minnow (Gobiocypris rarus) following perfluorooctanoic acid exposure. Comp. Biochem. Physiol. C. Toxicol. Pharmacol. 150, 57–64. Liu, C., Gin, K.Y.H., Chang, V.W.C., Goh, B.P.L., Reinhard, M., 2011. Novel perspectives on the bioaccumulation of PFCs - the concentration dependency. Environ. Sci. Technol. 45, 9758–9764. Liu, J., Lu, G., Ding, J., Zhang, Z., Wang, Y., 2014. Tissue distribution, bioconcentration, metabolism, and effects of erythromycin in crucian carp (Carassius auratus). Sci. Total Environ. 490, 914–920. Liu, H., Sheng, N., Zhang, W., Dai, J., 2015. Toxic effects of perfluorononanoic acid on the development of zebrafish (Danio rerio) embryos. J. Environ. Sci. (China) 32, 26–34. Liu, Z., Wang, Y., Zhu, Z., Yang, E., Feng, X., Fu, Z., Jin, Y., 2016. Atrazine and its main metabolites alter the locomotor activity of larval zebrafish (Danio rerio). Chemosphere 148, 163–170. Livak, K.J., Schmittgen, T.D., 2001. Analysis of relative gene expression data using realtime quantitative PCR and the 2−ΔΔCT method. Methods 25, 402–408. Lu, Z., Song, L., Zhao, Z., Ma, Y., Wang, J., Yang, H., Ma, H., Cai, M., Codling, G., Ebinghaus, R., Xie, Z., Giesy, J.P., 2015. Occurrence and trends in concentrations of perfluoroalkyl substances (PFASs) in surface waters of eastern China. Chemosphere 119, 820–827. Lushchak, V.I., 2011. Environmentally induced oxidative stress in aquatic animals. Aquat. Toxicol. 101, 13–30. Macaulay, L.J., Bailey, J.M., Levin, E.D., Stapleton, H.M., 2015. Persisting effects of a PBDE metabolite, 6-OH-BDE-47, on larval and juvenile zebrafish swimming behavior. Neurotoxicol. Teratol. 52, 119–126. Martin, J.W., Mabury, S.A., Solomon, K.R., Muir, D.C.G., 2003a. Bioconcentration and tissue distribution of perfluorinated aicds in rainbow trout (Oncorhynchus Mykiss). Environ. Toxicol. Chem. 22, 196–204. Martin, J.W., Mabury, S.A., Solomon, K.R., Muir, D.C.G., 2003b. Dietary accumulation of perfluorinated acids in juvenile rainbow trout (Oncorhynchus mykiss). Environ. Toxicol. Chem. 22, 189–195. Moody, D.E., Gibson, G.G., Grant, D.F., Magdalou, J., Rao, M.S., 1992. Peroxisome proliferators, a unique set of drug-metabolizing enzyme inducers: commentary on a symposium. Drug Metab. Dispos. 20, 779–791.

1406


Mortensen, A.S., Letcher, R.J., Cangialosi, M.V., Chu, S., Arukwe, A., 2011. Tissue bioaccumulation patterns, xenobiotic biotransformation and steroid hormone levels in Atlantic salmon (Salmo salar) fed a diet containing perfluoroactane sulfonic or perfluorooctane carboxylic acids. Chemosphere 83, 1035–1044. Nahrgang, J., Camus, L., Broms, F., Christiansen, J.S., Hop, H., 2010. Seasonal baseline levels of physiological and biochemical parameters in polar cod (Boreogadus saida): implications for environmental monitoring. Mar. Pollut. Bull. 60, 1336–1345. Ng, C.A., Hungerbühler, K., 2013. Bioconcentration of perfluorinated alkyl acids: how important is specific binding? Environ. Sci. Technol. 47, 7214–7223. Oakes, K.D., Sibley, P.K., Solomon, K.R., Mabury, S.A., van der Kraak, G.J., 2004. Impact of perfluorooctanoic acid on fathead minnow (Pimephales promelas) fatty acyl-CoA oxidase activity, circulating steroids, and reproduction in outdoor microcosms. Environ. Toxicol. Chem. 23, 1912–1919. Oakes, K.D., Sibley, P.K., Martin, J.W., Maclean, D.D., Solomon, K.R., 2005. Short-term exposures of fish to perfluorooctane sulfonate: acute effects on fatty acyl-CoA oxidase activity, oxidative stress, and circulating sex steroids. Environ. Toxicol. Chem. 24, 1172–1181. Oh, J.H., Moon, H.B., Choe, E.S., 2013. Alterations in differentially expressed genes after repeated exposure to perfluorooctanoate and perfluorooctanesulfonate in liver of Oryzias latipes. Arch. Environ. Contam. Toxicol. 64, 475–483. Pan, C.G., Zhao, J.L., Liu, Y.S., Zhang, Q.Q., Chen, Z.F., Lai, H.J., Peng, F.J., Liu, S.S., Ying, G.G., 2014. Bioaccumulation and risk assessment of per- and polyfluoroalkyl substances in wild freshwater fish from rivers in the Pearl River Delta region, South China. Ecotoxicol. Environ. Saf. 107, 192–199. Pessoa, P.C., Luchmann, K.H., Ribeiro, A.B., Veras, M.M., Correa, J.R., Nogueira, A.J., Bainy, A.C., Carvalho, P.S., 2011. Cholinesterase inhibition and behavioral toxicity of carbofuran on Oreochromis niloticus early life stages. Aquat. Toxicol. 105, 312–320. Ren, Q., Li, M., Yuan, L., Song, M., Xing, X., Shi, G., Meng, F., Wang, R., 2016. Acute ammonia toxicity in crucian carp Carassius auratus and effects of taurine on hyperammonemia. Comp. Biochem. Physiol. C. Toxicol. Pharmacol. 190, 9–14. Renzi, M., 2013. Perfluorinated organic compounds and polybrominated diphenyl ethers compounds – levels and toxicity in aquatic environments: a review. In: Rashed, M.N. (Ed.), Organic Pollutants: Monitoring, Risk and Treatment. IntechOpen, pp. 111–138. Richendrfer, H., Creton, R., Colwill, R.M., 2014. The embryonic zebrafish as a model system to study the effects of environmental toxicants on behavior. In: Lessman, C.A., Carver, E.A. (Eds.), Zebrafish. Nova Science Publishers, New York, pp. 245–264. Rotondo, J.C., Giari, L., Guerranti, C., Tognon, M., Castaldelli, G., Fano, E.A., Martini, F., 2018. Environmental doses of perfluorooctanoic acid change the expression of genes in target tissues of common carp. Environ. Toxicol. Chem. 37, 942–948. Rueda-Jasso, R., Conceição, L.E.C., Dias, J., De Coen, W., Gomes, E., Rees, J.F., Soares, F., Dinis, M.T., Sorgeloos, P., 2004. Effect of dietary non-protein energy levels on condition and oxidative status of Senegalese sole (Solea senegalensis) juveniles. Aquaculture 231, 417–433. Skutlarek, D., Exner, M., Färber, H., 2006. Perfluorinated surfactants in surface and drinking waters. Environ. Sci. Pollut. Res. 13, 299–307. Smith, R.A., Hartley, R.C., Cocheme, H.M., Murphy, M.P., 2012. Mitochondrial pharmacology. Trends Pharmacol. Sci. 33, 341–352. Sole, M., Fortuny, A., Mananos, E., 2014. Effects of selected xenobiotics on hepatic and plasmatic biomarkers in juveniles of Solea senegalensis. Environ. Res. 135, 227–235. Stahl, T., Mattern, D., Brunn, H., 2011. Toxicology of perfluorinated compounds. Environ. Sci. Eur. 23, 38. Ulhaq, M., 2013. The Toxicity of Perfluoroalkyl Acids in Zebrafish (Danio rerio). Department of Biomedical Sciences and Veterinary Public Health. Swedish University of Agricultural Sciences, Uppsala.

Ulhaq, M., Sundstrom, M., Larsson, P., Gabrielsson, J., Bergman, A., Norrgren, L., Orn, S., 2015. Tissue uptake, distribution and elimination of (14)C-PFOA in zebrafish (Danio rerio). Aquat. Toxicol. 163, 148–157. UNEP, 2009. SC-4/17: listing of perfluorooctane sulfonic acid, its salts and perfluorooctane sulfonyl fluoride. In: POPRC (Ed.), The Conference of the Parties, Geneva. van Boxtel, A.L., Kamstra, J.H., Cenijn, P.H., Pieterse, B., Wargna, J.M., Antink, M., Krab, K., van der Burg, B., Marsh, G., Brouwer, A., Legler, J., 2008. Microarray analysis reveals a mechanism of phenolic polybrominated diphenylether toxicity in zebrafish. Environ. Sci. Technol. 42, 1773–1779. Wang, P., Lu, Y., Wang, T., Fu, Y., Zhu, Z., Liu, S., Xie, S., Xiao, Y., Giesy, J.P., 2014. Occurrence and transport of 17 perfluoroalkyl acids in 12 coastal rivers in south Bohai coastal region of China with concentrated fluoropolymer facilities. Environ. Pollut. 190, 115–122. Wang, X., Halsall, C., Codling, G., Xie, Z., Xu, B., Zhao, Z., Xue, Y., Ebinghaus, R., Jones, K.C., 2014. Accumulation of perfluoroalkyl compounds in Tibetan mountain snow: temporal patterns from 1980 to 2010. Environ. Sci. Technol. 48, 173–181. Wang, T., Wang, P., Meng, J., Liu, S., Lu, Y., Khim, J.S., Giesy, J.P., 2015. A review of sources, multimedia distribution and health risks of perfluoroalkyl acids (PFAAs) in China. Chemosphere 129, 87–99. Wang, Z., Cousins, I.T., Scheringer, M., Hungerbuehler, K., 2015. Hazard assessment of fluorinated alternatives to long-chain perfluoroalkyl acids (PFAAs) and their precursors: status quo, ongoing challenges and possible solutions. Environ. Int. 75, 172–179. Wei, Y., Chen, L.L., Wang, D., Zhang, H., Wang, J., Dai, J., 2008a. Proteomic analysis of hepatic protein profiles in rare minnow (Gobiocypris rarus) exposed to perfluorooctanoic acid. J. Proteome Res. 7, 1729–1739. Wei, Y., Liu, Y., Wang, J., Tao, Y., Dai, J., 2008b. Toxicogenomic analysis of the hepatic effects of perfluorooctanoic acid on rare minnows (Gobiocypris rarus). Toxicol. Appl. Pharmacol. 226, 285–297. Xia, J., Cao, Z., Peng, J., Fu, S., Fu, C., 2014. The use of spontaneous behavior, swimming performances and metabolic rate to evaluate toxicity of PFOS on topmouth gudgeon Pseudorasbora parva. Acta Ecol. Sin. 34, 284–289. Xie, Z., Lu, G., Hou, K., Qin, D., Yan, Z., Chen, W., 2016. Bioconcentration, metabolism and effects of diphenhydramine on behavioral and biochemical markers in crucian carp (Carassius auratus). Sci. Total Environ. 544, 400–409. Yang, J.H., 2010. Perfluorooctanoic acid induces peroxisomal fatty acid oxidation and cytokine expression in the liver of male Japanese medaka (Oryzias latipes). Chemosphere 81, 548–552. Yang, S., Liu, S., Ren, Z., Jiao, X., Qin, S., 2014. Induction of oxidative stress and related transcriptional effects of perfluorononanoic acid using an in vivo assessment. Comp. Biochem. Physiol. C. Toxicol. Pharmacol. 160, 60–65. Yang, Y., Qi, S., Wang, D., Wang, K., Zhu, L., Chai, T., Wang, C., 2016. Toxic effects of thifluzamide on zebrafish (Danio rerio). J. Hazard. Mater. 307, 127–136. Yeh, A., 2017. Modulation of Mitochondrial Function as an Indicator for Sublethal Effects of Contaminants of Emerging Concern. Department of Environmental and Occupational Health Sciences. University of Washington, Seattle, p. 117. Zhang, W., Liu, Y., Zhang, H., Dai, J., 2012a. Proteomic analysis of male zebrafish livers chronically exposed to perfluorononanoic acid. Environ. Int. 42, 20–30. Zhang, W., Zhang, Y., Zhang, H., Wang, J., Cui, R., Dai, J., 2012b. Sex differences in transcriptional expression of FABPs in zebrafish liver after chronic perfluorononanoic acid exposure. Environ. Sci. Technol. 46, 5175–5182.