Aquatic Toxicology 163 (2015) 148–157

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Tissue uptake, distribution and elimination of 14 C-PFOA in zebrafish (Danio rerio) Mazhar Ulhaq a,1 , Maria Sundström b,2 , Pia Larsson a , Johan Gabrielsson a , Åke Bergman b,3 , Leif Norrgren a , Stefan Örn a,∗ a b

Department of Biomedical Sciences and Veterinary Public Health, Swedish University of Agricultural Sciences, SE-750 07 Uppsala, Sweden Environmental Chemistry Unit, Department of Materials and Environmental Chemistry, Stockholm University, SE-106 91 Stockholm, Sweden

a r t i c l e

i n f o

Article history: Received 9 December 2014 Received in revised form 31 March 2015 Accepted 2 April 2015 Available online 3 April 2015 Keywords: Zebrafish PFOA Perfluorinated chemicals Radiolabeled Kinetics

a b s t r a c t Perfluorooctanoic acid (PFOA) is a long-chain perfluorinated chemical that has been shown to be non-degradable and persistent in the environment. Laboratory studies on bioconcentration and compound-specific tissue distribution in fish can be valuable for prediction of the persistence and environmental effects of the chemicals. In the present study male and female zebrafish (Danio rerio) were continuously exposed to 10 ␮g/L of radiolabeled perfluorooctanoic acid (14 C-PFOA) for 40 days, after which the exposed fish were transferred to fresh clean water for another 80 days wash-out period. At defined periodic intervals during the uptake and wash-out, fish were sampled for liquid scintillation counting and whole body autoradiography to profile the bioconcentration and tissue distribution of PFOA. The steady-state concentration of 14 C-PFOA in the zebrafish was reached within 20–30 days of exposure. The concentration-time course of 14 C-PFOA displayed a bi-exponential decline during washout, with a terminal half-life of approximately 13–14 days. At steady-state the bioconcentration of 14 C-PFOA into whole-body fish was approximately 20–30 times greater than that of the exposure concentration, with no differences between females and males. The bioconcentration factors for liver and intestine were approximately 100-fold of the exposure medium, while in brain, ovary and gall bladder the accumulation factors were in the range 15–20. Whole-body autoradiograms confirmed the highest labeling of PFOA in bile and intestines, which implies enterohepatic circulation of PFOA. The 14 C-PFOA was also observed in maturing vitellogenic oocytes, suggesting chemical accumulation via yolk proteins into oocytes with plausible risk for adverse effects on early embryonic development and offspring health. The bioconcentration at several 14 C-PFOA exposure concentrations were also investigated (0.3–30 ␮g/L). This showed that bioconcentration increased linearly with tank exposure in the present in vivo model under steady-state conditions. From this model tissue concentrations of PFOA can be predicted when the external exposure level is known. The present study has generated experimental data on PFOA kinetics in zebrafish that can be valuable for aquatic environmental risk assessment. © 2015 Elsevier B.V. All rights reserved.

1. Introduction

∗ Corresponding author. Tel.: +46 18 671178. E-mail address: [email protected] (S. Örn). 1 Current affiliation: Department of Biomedical Sciences, PMAS Arid Agriculture University, PK-46300 Rawalpindi, Pakistan. 2 Current affiliation: Department of Applied Environmental Science (ITM), Stockholm University, SE-10691 Stockholm, Sweden. 3 Current affiliation: Swedish Toxicology Sciences Research Center (Swetox), Forskargatan 20, SE-15136 Södertälje, Sweden. http://dx.doi.org/10.1016/j.aquatox.2015.04.003 0166-445X/© 2015 Elsevier B.V. All rights reserved.

Perfluoroalkyl acids (PFAAs) are man-made industrial chemicals that have been in use since the 1940s. PFAAs form a family of synthetic fully fluorinated hydrocarbons with a functional group of carboxylate, sulfonate or phosfonate. PFAAs have unique properties of strongly reducing surface tension and inertness to thermal and chemical degradation These properties have made them extensively used in various industrial applications and consumer products, such as surfactants, emulsifiers, fire-fighting foams, stain, dirt and oil-resistant coatings for carpets, leather, fabrics, food packaging paper products, waxes and car polishes (Paul et al., 2009).

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PFAAs have been shown to be protein-binding and nonbiodegradable enzymatically, making them persistent and bioaccumulative in humans and wildlife, including animals in remote locations such as polar bears (Calafat et al., 2007; Giesy and Kannan, 2001; Persson et al., 2013; Yamashita et al., 2008; Han et al., 2005). For water-living organisms, discharges from municipal sewage water treatment plants are assumed to be significant sources for contamination of aquatic environments. Several studies indicate the presence of these chemicals in wastewater and their final release into sludge and natural waters (Olofsson et al., 2013; Pan et al., 2011; Yeung et al., 2009). The toxicity of PFAAs is generally determined by the carbon chain length of the individual molecule and the functional group attached (Buhrke et al., 2013; Hagenaars et al., 2011; Matsubara et al., 2006; Reistad et al., 2013; Ulhaq et al., 2013a,b; Wolf et al., 2008; Zheng et al., 2012). In fish toxicity studies, developmental malformations, physiological disturbances and impaired larval behavior have been associated with exposure to different PFAAs (Hagenaars et al., 2011; Ulhaq et al., 2013a,b; Zheng et al., 2012). Perfluorooctanoic acid (PFOA) is one of the most widely detected PFAAs in humans and environmental samples (Domingo et al., 2012; Kannan et al., 2005; Miege et al., 2012; Rudel et al., 2011). It has been produced in large quantities, and can also be formed as a metabolite from other perfluorinated chemicals. The exposure pathways of PFOA and related PFAAs to humans are the subject of investigation. However, drinking water (Murray et al., 2010), and the consumption of fish and meat (Falandysz et al., 2006; Guruge et al., 2005), are considered to be major contributors of PFAAs to human exposure. Although the production of PFOA has decreased in recent years, the chemical will remain a health and risk concern due to its presence in household and commercial products manufactured before the start of the stewardship program (Betts, 2007). Unlike many other chemicals, PFOA is not primarily accumulated in the adipose tissue but rather in the hepatic tissue and blood. PFOA binds to blood serum proteins and undergoes enterohepatic circulation (Johnson et al., 1984; Jones et al., 2003). Several experimental studies with mammals indicate species- and genderspecific kinetic behaviors of PFOA (Lee and Schultz, 2010; Ohmori et al., 2003; Weaver et al., 2010). Toxicokinetic studies show that PFOA is readily absorbed after oral administration, slowly distributed and then slowly eliminated. PFOA appears not to be biotransformed in rats (Johnson et al., 1984; Vanden Heuvel et al., 1991). Studies also indicate large species related differences in the half-life of PFOA (Burris et al., 2002; Hanhijarvi et al., 1982; Lou et al., 2009; Ohmori et al., 2003; Yoo et al., 2009). Urinary excretion is reported to be the most important elimination pathway in rodents. Large gender-related differences in urinary elimination have been reported with shorter plasma half-lives in female rats compared to males (Loccisano et al., 2011; Vanden Heuvel et al., 1991). Gender differences in the elimination of PFOA have also been observed in fish species like fathead minnows (Pimephales promelas) and Nile tilapia (Oreochromis niloticus) (Han et al., 2011; Lee and Schultz, 2010). The objective of this study was to apply an equilibrium model to establish scenarios similar to natural environmental conditions, and then to examine the uptake, distribution and elimination of PFOA in zebrafish (Danio rerio). An equilibrium model can mimic chronic environmental exposure of the test compound and thereby estimate the accumulation potential to be used in risk assessment. Whole Body Autoradiograhy (WBA) gives a rapid and precise visualization of the spatial distribution of radiolabeled chemicals in organs and tissues of an intact animal (Ullberg 1954). Thereby WBA is complementary to biokinetic quantification based on scintillation of dissected organs in order to describe the fate of chemicals. Moreover, determination of PFOA accumulation in target tissues can

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lead to understanding of the physiological effects under realistic environmental conditions. 2. Material and methods 2.1. Chemicals Hydrogen peroxide (30% in water) and ethyl 3-aminobenzoate methanesulfonic acid (MS-222), an anesthetic agent, were purchased from Sigma–Aldrich (Germany). The solubilization reagent Soluene® -350, Hionic-FlourTM Scintillation cocktail and 14 C carbon dioxide were purchased from PerkinElmer (PerkinElmer Life and Analytical Sciences, Boston, USA). Perfluoroheptyl iodide came from TCI Europe (Tokyo, Japan). All other chemicals used were obtained from regular commercial sources and of pro-analysis quality. 2.2. Synthesis of 14 C-PFOA The radiolabeled perfluorooctanoic acid (14 C-PFOA) was synthesized as described in detail by Sundström (2012). The synthesis was developed from a previous method published by Shtarov and Howell (2005) accomplished by introducing a solution of ethyl magnesium bromide (3 M, 1.0 mmol, 0.33 mL) at −50 ◦ C into the reaction flask containing diethyl ether (dried, 10 mL) and perfluoroheptyl iodide (0.93 mmol, 0.22 mL) under an inert atmosphere. The Grignard reagent was stirred at −50 ◦ C for one hour before the temperature was decreased to approximately −100 ◦ C. Tetrahydrofuran (10 mL) was added to increase the solubility of carbon dioxide in the reaction mixture. The 14 C-carbon dioxide (50 mCi, 59 mCi/mmol) was delivered in a break-seal flask which was opened and transferred under vacuum according to the supplier’s guidelines. After transfer of the carbon dioxide into the reaction flask the temperature was allowed to increase to −78 ◦ C during four hours and thereafter slowly to room temperature overnight. The reaction was cooled to −30 ◦ C and carefully quenched with water (0.5 mL) and sulfuric acid (10 M, 1 mL). The ether phase was transferred and collected in a specialized designed flanged evaporation beaker. This beaker prevented to some degree the formation of micelles (excessive foaming) during evaporation. The raw product was extracted with diethyl ether (4 × 5 mL) that gave, after evaporation of the solvent, an orange-colored oil to which potassium hydroxide (2 M, 2 mL) was added. The formed colorless and gel-like mass was extracted with n-pentane (3 × 8 mL). Hydrochloric acid (37%, 1 mL) was carefully added to the water phase before extraction with dichloromethane (8 × 6 mL) followed by evaporation on a rotary evaporator. The synthesis yielded 20 mCi of 14 C-PFOA with a specific activity of 59 mCi/mmol. The radiochemical purity was 97% as determined by LC–MS analysis. 2.3. Animals Zebrafish were purchased from a local fish shop and then acclimatized for three weeks prior to the studies. Throughout the experiments the fish were held in charcoal filtered oxygenated tap water at a temperature of 25 ± 2 ◦ C, pH 8.0 ± 0.3, hardness 8.4 ± 0.6 dH, conductivity 48.2 ± 1.4 mS/m and at a 14-h light cycle. The fish were fed two times daily with flakes (Nutrafin), frozen red blood worms and Artemia. Studies performed were approved by the local Ethical Committee for Animal Experimentation (Dnr C309/11). 2.4. Experimental overview The zebrafish were divided into groups with phenotypic males and females kept in separate tanks for use in two different

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experimental studies (Experiments 1 and 2, see description below). In Experiment 1, males and females were exposed to a single concentration of 10 ␮g/L of PFOA to study uptake, distribution and elimination. Fish were exposed for 40 days, which was followed by a washout period of 80 days. During the experimental period, fish were regularly sampled for Liquid Scintillation Counting (LSC) and whole body autoradiography (WBA). In Experiment 2, males and females were exposed to five different exposure concentrations of PFOA (0.3–30 ␮g/L) and sampled after 40 days for LSC for studies of bioconcentration in organs and whole-body fish. 2.4.1. Experiment 1 Male and female zebrafish were exposed separately in 50-L tanks to a nominal exposure concentration of PFOA at 10 ␮g/L (1.4 ␮Ci 14 C-PFOA/L). Partial water changes of 2–3% were done twice per week to remove debris. The 14 C-PFOA activity in the exposure medium was measured by LSC daily, with deviations in concentrations being corrected accordingly, thus keeping the nominal concentration at ≥90% throughout the exposure period of the experiment. During the exposure period, fish were sampled for LSC after 1, 5, 10, 15, 20, 30 and 40 days. At termination of exposure, the remaining fish were transferred to 50-L water tanks without PFOA. Here, partial water changes of 90% were done twice per week. Fish were subsampled after 1, 3, 8, 16, 30, 55 and 80 days during the washout period. Five male and five female zebrafish were sampled at each time point for LSC. In addition, three males and females were sampled for WBA after 1, 5, 10 and 30 days during the exposure period, and after 4, 8, 16 and 30 during the washout period. 2.4.1.1. Liquid scintillation counting. The radioactivity of the water samples was analyzed regularly. Water samples (1 mL) were mixed with 10 mL Hionic-FlourTM scintillation counting cocktail, and analyzed by Tri-carb 1900CA Liquid Scintillation Analyzer (Packard). The sampled zebrafish were immersed in clean water, euthanized in 100 mg/L MS-222 sodium bicarbonate buffered (equal w/w) aqueous solution, and thereafter weighed. Each fish was chopped to a maximum 40 mg of tissue, which was then transferred into 20 mL glass scintillation vials (VWR, Sweden). The tissues were solubilized in Soluene® – 350 at 60 ◦ C for 24 h, and agitated intermittently to ensure complete solubilization. The samples were then allowed to cool to room temperature. Some samples were bleached with 30% hydrogen peroxide and left at room temperature for 1 h. Ten milliliter of scintillation cocktail Hionic-FlourTM was added to all samples. Before LSC, the samples were placed in ambient light and temperature for 1 h to minimize background counts. The tissue contents of 14 C-PFOA were determined by LSC, and based on these results the total concentration of PFOA was calculated from the specific activity of 14 C-PFOA in the water. The activities of the individual fish samples were summed, and adjusted for the weight of the fish, to determine the concentration of PFOA. 2.4.1.2. Whole-body autoradiography. The sampled fish were first immersed for 30 s in clean water and euthanized in MS-222. The fish were then mounted in aqueous carboxymethyl cellulose (CMC) gel, and frozen in a bath of hexane cooled with dry ice. The frozen blocks were processed further for WBA as described by Larsson and Ullberg (1981). A series of whole-body sagittal 20 ␮m thick sections was taken at different levels, and collected on tape and freeze-dried. To study the distribution of non-extractable radioactivity every other freeze-dried tissue section was extracted with 5% trichloroacetic acid, 50% ethanol, 99.5% ethanol and heptane for 1 min, and then rinsed with tap water for 5 min. The sections were dried and pressed against X-ray film and stored at −20 ◦ C for exposure. The exposed films were then developed between 8 weeks and up to 9 months. To be able to simultaneously illustrate the uptake during the exposure and elimination period, the

radioluminographic technique was used. This technique is more sensitive than traditional exposure to X-ray film, and makes it possible to detect a wide range of radioactivity levels. Thus, sections from 1, 5, 10 and 30 days of exposure, and sections from days 8, 16 and 30 of the washout period, were exposed to phosphor imaging plates (Storage Phosphorous Screen). Radioluminographic images were then obtained by developing the plate using the Packard Cyclone® Plus Storage System (PerkinElmer, IL, USA). 2.4.1.3. Kinetic assessment. Data were collected over the 120-day period, and total body exposure was assessed by means of concentration-time and amount-time data. The noncompartmental approach was used as a means to assess the major determinants of exposure, namely body clearance (removal) and uptake-rate (input), and based on the assumption that PFOA is chemically and metabolically stable during the observational period (Ophaug and Singer, 1980; Sundström et al., 2012; Vanden Heuvel et al., 1991; Ylinen et al., 1989). Since blood or plasma was not available for the estimation of plasma clearance, we applied the concepts of total body clearance based on the total body tissue concentration (Gabrielsson and Weiner, 2010). Body clearance Clb was estimated from the amount of PFOA at steady-state Ass and the post-exposure (washout period) area from t* (end of tank exposure period to PFOA) to infinity AUCt∗-∞ derived by the trapezoidal method. The relationship between amount of PFOA at steady-state, total body clearance and the post-exposure area under the exposure curve is given by Eq. (1): Ass = Clb × AUC∞ t∗

(1)

This can be re-arranged to yield body clearance (Eq. (2)). Ass is estimated from the total body concentration (whole fish) of 14 CPFOA at steady-state multiplied by their body weight(s). Clb =

ASS AUC∞ t∗

(2)

The total body clearance (based on total body concentration rather than plasma PFOA concentration) can be related to plasma clearance (Eq. (3)) provided plasma concentration is obtained from the specific animal model (e.g., zebrafish). Clp =

Cb × Clb Cp

(3)

Here, Clp , Cp , Clb and Cb are the plasma clearance, plasma concentration, body clearance and total body concentration of PFOA, respectively. The mean residence time MRT of 14 C-PFOA in zebrafish, which denotes the time an average molecule of PFOA remains in the organism (Eq. (4)), can be calculated from the areas of the first- and zero-moment curves, and then correction for the 40-day exposure period (Texposure ).



MRT =

AUMC0∞ AUC∞ 0



Texposure 2



(4)

Where AUMC0 - ∞ denotes the area under the first-moment curve, AUC0-∞ the area under the zero moment curve and Texposure the length of the external (tank) exposure period (Benet, 2010; Gabrielsson and Weiner 2010). The effective half-life t1/2 (e) is then derived from MRT by: t 1/2 (e) = ln(2) × MRT

(5)

The effective half-life is a weighted half-life related to the turnover of the amount of PFOA in the body, rather than the terminal half-life of test compound commonly obtained from the terminal phase of the plasma concentration-time curve. Three to four times the effective half-life gives an indication of the time to 90% of steady-state in the body during a constant exposure to the

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Fig. 1. (A) Schematic representation of 14 C-PFOA exposure concentration versus zebrafish internal concentration after exposure. (B) Fish concentration is proportional to the uptake/elimination of 14 C-PFOA by means of body clearance.

test compound. Fig. 1 shows schematically the respective parameters of Eq. (6). At steady state the rate of uptake equals the rate of elimination (Eq. (6)), which gives: Rateofuptake = Cb × Clb = Rateofelimination

(6)

Since Cb is measured and Clb is given by Eq. (1) and (2), the rate of uptake can be predicted. Provided body clearance Clb is constant within the experimental range, rate of uptake can be calculated when body exposure Cb is known. The terminal half-life of PFOA in male and female zebrafish was estimated directly from loglinear regression of the terminal portion of the wash-out kinetics. At the end of the exposure, the chemical profiles in the whole fish homogenate were described by the accumulation ratio (AR). AR is similar to bioconcentration factor where the numerator of the fraction is the concentration of PFOA calculated per gram wet weight of whole fish, and the denominator is exposure concentration per milliliter of water. 2.4.2. Experiment 2 Male and female zebrafish were exposed separately in 5-L tanks to nominal exposure concentrations of PFOA at 0.3, 1, 3, 10 and 30 ␮g/L (0.04, 0.14, 0.42, 1.4 and 4.2 ␮Ci 14 C-PFOA/L, respectively). Partial water changes of 2–3% were done twice per week to remove debris. The 14 C-PFOA activity in the exposure medium was measured by LSC daily, with deviations in concentrations being corrected accordingly, thus keeping the nominal concentration at ≥90% throughout the exposure period of the experiment. After 40 days of exposure, five males and five females from each exposure group were sampled for LSC. In addition, three males and females were sampled and dissected for determination of PFOA in different organs (i.e., liver, intestine, brain, ovary). By measuring the total body concentration of PFOA on day 40 of exposure, a relationship could be established by plotting water concentrations to fish concentrations (Fig. 1a). Fish concentration combined with the estimated body clearance of 14 C-PFOA then gives the uptake or elimination rate at steady-state (Fig. 1b). This conceptual diagram demonstrates the usefulness of the clearance concept that is shown in Fig. 2. The bioconcentration factor (BCF) for each fish was calculated using the equation: BCFss =

Cfss Cwss

(7)

Here BCFss is the bioconcentration factor at steady state; Cfss is the concentration of 14 C-PFOA in whole fish at steady state; and Cwss is

the concentration of the test substance in the water at steady state. The BCF is defined as the volume of exposure water that is depleted of PFOA by the fish within the exposure period (mL/g body weight). 3. Results In both Experiments 1 and 2, the measured concentrations of PFOA in water were close to the nominal concentrations (≥90%) throughout the exposure period. In Experiment 1, the mean PFOA concentrations during the exposure period were 9.58 ± 0.48 ␮g/L and 9.43 ± 0.45 ␮g/L, in male and female tanks, respectively. In Experiment 2, the measured PFOA concentrations over the exposure period were 0.31 ± 0.01, 1.10 ± 0.11, 3.14 ± 0.19, 10.27 ± 0.69 and 30.25 ± 0.76 ␮g/L in male tanks, and 0.30 ± 0.01, 1.09 ± 0.13, 3.14 ± 0.12, 10.45 ± 0.75 and 30.61 ± 0.83 ␮g/L in female tanks. PFOA were not detected in control tanks in either experiments. No mortalities were observed in either experiments. In Experiment 1, both male and female zebrafish absorbed PFOA relatively rapidly from the water. Already after 5 days of exposure to 10 ␮g/L, the mean whole-body concentrations were 125 ± 28 ng/g in males and 98 ± 15 ng/g in females, respectively (Fig. 3). Clear increases in concentrations were then measured up to days 15–20. The equilibrium concentrations occurred between 20 and 30 days of exposure. After 40 days of exposure, the mean concentrations of PFOA were 489 ± 107 ng/g in males and in 239 ± 61 ng/g in females, respectively (Fig. 3). The non-compartmental calculations of body clearance (Clb ), effective (T½ (e)) and terminal (T½ (z)) half-lives estimated from uptake/washout data, and the accumulation factor in male and female zebrafish at steady-state shows that uptake and disposition kinetics of PFOA demonstrated an average total body clearance of 50 mL/day, with terminal and effective half-lives of 13–14 and 7–8 days, respectively (Table 1). No significant differences were recorded between males and females with respect to the disposition kinetic parameters. The whole-body radioluminograms of fish sampled during uptake and wash out illustrate the most important tissues for accumulation and retention of 14 C-PFOA (Fig. 8). The distribution of 14 C-PFOA in the whole-body autoradiograms are representative for the fish sampled at each given exposure time and concentration group. After 1 day of exposure, the fish had accumulated a substantial amount of the radiolabeled compound (Fig. 6). Wholebody autoradiograms of fish exposed for 1, 5 and 10 days showed a similar progressive distribution pattern of 14 C-PFOA (Fig. 6A and B). The highest amounts of radioactivity were concentrated in bile

Table 1 Selected disposition parameters of 14 C-PFOA in zebrafish. Clb = body clearance; MRT = mean residence time; T1/2 (e) = effective half-life; T1/2 (z) = terminal half-life; SS = steady state. Gender Female Male

Clearance Clb (mL/day) −3

42 × 10 56 × 10−3

MRT (days)

T1/2 (e) (days)

T1/2 (z) (days)

Accumulation factor at SS

11 10

7.9 7.2

13 14

20–30 20–30

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Fig. 2. Schematic diagram of the method used for estimation of body clearance (Eqs. (1) and (2)) of 14 C-PFOA. Mean 14 C-PFOA concentration-time data are presented on a semi-logarithmic plot.

and in the intestines. The structures perfused by blood also showed a considerable labeling, e.g., the heart, gills and blood vessels along the vertebral column (Fig. 6). A relatively high labeling was also observed in the choroid of the eye, the olfactory rosette, and in the head and trunk kidney (Fig. 6). The labeling in the liver was relatively low and uniform. In addition, the superficial parts of the skin and fins also showed distinct labeling. A diffuse and weak labeling could also be seen in other tissues. In exposed female fish, radioactive labeling was observed in the oocytes at all sampling times. Initially, only the outer membranes of the oocytes were labeled, and with the advancement of exposure a moderate labeling was developed throughout vitellogenic oocytes, while other immature oocytes still lacked labeling (Fig. 7A and B). The spleen showed a weak heterogeneous labeling. Skeletal muscles and brain showed faint or no labeling. However, certain structures within these tissues showed a distinct labeling, probably corresponding to blood vessels. Extracted sections showed no 14 C-PFOA labeling. In the autoradiograms of fish sampled on days 8, 16 and 30 of the washout periods, the radioactivity progressively declined in almost all tissues. The decrease in the radioactive intensity was slower in the liver, intestines, olfactory rosette and the oocytes. In the bile, 14 CPFOA declined very slowly. On day 30 of wash-out only the bile still showed a very high intensity of radioactive labeling, whereas other tissues had none or only weak labeling (Fig. 8). In Experiment 2, the steady-state actual water concentration was plotted against the tissue concentration of 14 C-PFOA after 40 days of exposure. An apparently linear relationship was obtained between exposure concentration and whole-body

tissue homogenates concentrations (Fig. 4). The PFOA concentrations measured in whole-body samples of fish exposed to 0.3–30 ␮g/L PFOA ranged between 10 ± 3 and 1300 ± 600 ng/g in males and 10 ± 5–670 ± 80 ng/g in females after 40 days of exposure. No statistical differences in bioconcentration between males and females were observed in either whole-body homogenates or the individual organ concentrations (Fig. 5). 4. Discussion The present study utilized a two-step experimental design for assessing the uptake and disposition of 14 C-PFOA in zebrafish. The uptake of PFOA was relatively rapid with equilibrium concentrations reached after 20–30 days of exposure. The whole-body bioconcentration factor (BCF) for PFOA in zebrafish was 20–30, which is in agreement with rainbow trout studies where BCF ranged between 4 and 27 (Martin et al., 2003a). In comparison, BCF for PFOS in fish is generally reported to be at least 1000 times higher than PFOA (Kannan et al., 2005). The elimination study showed that the concentration-time course of PFOA displayed a bi-exponential decline during the washout phase, with an initial 7–8 days effective half-life followed by a slower terminal half-life of approximately 13–14 days. This suggests that total body exposure to PFOA mimics multi-compartment disposition, which means that a substantial amount of PFOA is allocated to rapidly equilibrated tissue deposits as well as more slowly equilibrated deposits. Based on the 40days exposure when fish were sampled regularly a total body clearance of 50 mL/day was estimated. The effective half-life of

Fig. 3. Semi-logarithmic plot of all individual concentration-time data points of 14 C-PFOA in zebrafish. Each point represents one individual male (filled squares) or female zebrafish (open circles). The nominal tank exposure was 10 ␮g/L.

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Fig. 4. Log–log plot of the whole-body homogenate concentrations in individual fish on day 40 of exposure versus actual 14 C-PFOA exposure concentrations in tanks. Five groups of male and female zebrafish were exposed to 0.3, 1.0, 3.0, 10 and 30 ␮g L−1 of 14 C-PFOA. Each point represents one individual male (filled squares) or female (open circles). The straight lines denote the linear regression predictions for males and females.

Fig. 5. Log–log plot of actual external 14 C-PFOA exposure versus mean internal exposure in liver, intestine, gonads and brain tissues under steady state conditions. Groups of male and female zebrafish were exposed to 0.3, 1.0, 3.0, 10 and 30 ␮g/L of 14 C-PFOA. Male animals (solid curves) and female animals (dashed curves). Note the 10–20 fold accumulation factors relative to tank exposure in male and female brain tissue, and the almost 100-fold accumulation factor in liver and intestinal tissue.

7 days indicates that the 40-days (>5 half-lives) exposure period is adequate to mimic chronic exposure to 14 C-PFOA. In the bioconcentration study, which covered a 100-fold external exposure range (0.3–30 ␮g/L) a variation across different tissues with respect

Fig. 6. (A) Whole-body autoradiogram of a female zebrafish exposed to 14 C-PFOA for 24 h. (B) Histological sections of the same fish as in A showing organs. G – gills, H – heart, L – liver, Gb – gallbladder, E – eye, K – kidney, Oo – olfactory organ, S – spinal cord, O – ovary, Sb – swimbladder.

to accumulation of 14 C-PFOA was measured. The accumulation of 14 C-PFOA into individual tissues ranged from 20 (brain) to 100-fold (liver) times higher than the external exposure. These results are in agreement with previous studies of PFOA uptake in both zebrafish, Daphnia magna, green mussel, earthworm, common carp and rainbow trout (Dai et al., 2013; Hagenaars et al., 2013; Liu et al., 2011; Martin et al., 2003a; Ng and Hungerbuhler, 2013). In the present study no chemical measurements of background levels of nonradioactive PFOA, or of other chemicals, in the fish or in the food were done. As for all studies, there is a risk that chemicals other than the studied can influence the results, e.g., by affecting physiological mechanisms involved in the uptake, distribution or elimination of the model chemical. There are large differences in reported biological half-lives of PFOA among vertebrates, e.g., 5.5 h in rabbits, 4–6 days in rats, 3–13 days in fish, 12 days in mice, 21 days in monkeys and 20–30 days dogs (Butenhoff et al., 2004; Han et al., 2011; Hanhijarvi et al., 1988, 1982; Hundley et al., 2006; Lau et al., 2007; Lee and Schultz, 2010; Consoer et al., 2014; Martin et al., 2003b, 2004). In fish, bioaccumulation and tissue distribution of PFOA have been investigated experimentally in different species, e.g., rainbow trout (Oncorhynchus mykiss) (Consoer et al., 2014; Martin et al., 2003a,b), Chinese sturgeon (Acipensersinensis) (Peng et al., 2010), blackrock fish (Sebastes schlegeli) (Jeon et al., 2010) common carp (Cyprinus carpio L.) (Inoue et al., 2012) and fathead minnow (P. promelas) (Lee and Schultz, 2010). Also among fish species

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Fig. 7. (A) Ovary with oocytes (arrows) from a zebrafish exposed to washout showing labeling throughout whole oocytes (arrows).

14

C-PFOA for 1 day. Only the eggshell of oocytes are labeled. (B) Ovary from a zebrafish on day 8 of

differences in half-lives of PFOA has been measured. In the present study, the terminal half-life of PFOA in zebrafish (13 days) was equal that reported for rainbow trout (Consoer et al., 2014). However, in other studies with rainbow trout half-lives of 0.26 and 6 days were reported (Martin et al., 2003a,b; De Silva et al., 2009). Generally, elimination rate is influenced by body size, but also other anatomical and physiological factors can be of importance. In some species, the half-life of PFOA is also gender-related, for instance female rats excrete PFOA very quickly (0.08 day) compared with male rats (5.68 days) i.e., almost 70 times faster than males (Kudo et al., 2002). The opposite gender difference excretion pattern occurs in hamsters (Hundley et al., 2006). The phenomenon of gender difference in elimination of PFOA has also been shown in fish. Female fathead minnows (P. promelas) and Nile tilapia (O. niloticus) have shorter PFOA half-life (0.26, 0.014 day) compared to males (2.85, 0.058 day) (Han et al., 2011; Lee and Schultz, 2010). In the present study, zebrafish showed no overall apparent gender differences in the uptake and disposition of 14 C-PFOA, neither in specific tissues (liver, intestine, brain) nor in whole-body uptake over the 100-fold exposure range concentrations (0.3–30 ␮g/L). In contrast, Hagenaars et al. (2013) reported a gender difference in accumulation of PFOA at a 4 days short term exposure of zebrafish but not when applying a 28 days long-term exposure. The authors discussed that the lack in gender differences at the prolonged exposure might be caused by disturbances in sex steroid metabolism affecting renal excretion (Hagenaars et al., 2013) which has been observed also in fathead minnow and rats exposed to PFOA (Lee and Schultz, 2010; Vanden Heuvel et al., 1991; Oakes et al., 2004).

The labeling of 14 C in brain and skeletal muscles is probably related to blood vessels. Skeletal muscle is poorly perfused as compared to the highly perfused organs such as heart, gills, brain and spleen (Nichols et al., 2004). Highly perfused organs and the richly vascularized eye choroid showed a relatively higher labeling than that of the skeletal muscles in exposed fish. The tissue labeling other than excretory pathways (hepatobiliary and urinary) might be related to the relative vascularization and diffusion of radiolabeled chemical in the tissues. In 14 C-PFOA exposed zebrafish, labeling of the kidney was observed, though relatively less than intestine and bile. In exposed rainbow trout, the concentration of PFOA was greatest in kidneys, after blood, among all other organs (Martin et al., 2003b). Measurements as well as weak labeling of PFOA in the brain and olfactory bulb of the brain indicate these as target tissues for potential accumulation and effect. The uptake of metals through the olfactory epithelium has already been observed in rats and fish (Persson et al., 2003; Tjälve et al., 1995, 1996). The axonal transport of PFOA and other PFAAs has not, to our knowledge, been reported. The uptake and transport of environmental chemicals through the olfactory epithelium is an active route of access to the central nervous system, and may be related to neurological dysfunctions. Similar to our laboratory findings, PFAAs have also been measured in the brain tissues of wild fish (Shi et al., 2012). However, the brain data should be interpreted very carefully, because the complex vascular structures of the brain may contaminate the tissue, as blood vessels containing 14 C-PFOA may be mistakenly included while extracting the tissue. Direct uptake into brain regions is possible, as chemicals can cross the blood–brain barrier. This phenomenon has already been observed for PFAAs in freshwater farmed fish

Fig. 8. Zebrafish whole-body radioluminograms illustrating distribution of 14 C-PFOA on day 1, 5, 10, 30 of uptake and 8, 16, 30 of washout.

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(Shi et al., 2012) along with other chemicals such as metals (Yokel, 2006). The nervous system is a sensitive target for various chemicals and PFAAs have been linked as causative agents of mental disorders (Barkley, 1998; Brown et al., 2005; Hoffman et al., 2010). PFAAs are considered to be developmental neurotoxicants, and effects on spontaneous behavior, habituation capability, learning and memory later at adulthood have been observed in prenatally or neonatal exposed mice (Johansson et al., 2008; Onishchenko et al., 2011). In epidemiological studies, correlations between PFAA exposure levels and ADHD (attention deficit hyperactivity disorder) in children have been reported (Fei et al., 2008; Hoffman et al., 2010). In laboratory studies with zebrafish, behavioral alterations including spontaneous hyperactivity has been observed after exposure to PFOS (Spulber et al., 2014). The reason for the high concentrations of PFOA in the gall bladder, liver and also in the intestinal tissues is probably due to enterohepatic circulation of this chemical. This mechanism has been observed also in laboratory studies with rainbow trout (Martin et al., 2003b) and in rodents (Johnson et al., 1984; Kudo and Kawashima, 2003). Enterohepatic recirculation increases the risk for hepatotoxicity, and in fact most PFAAs show similar toxicity profiles with liver as the main target organ (Lau et al., 2007). No metabolism has been shown for PFOA and it is considered being excreted intact without forming any metabolites or conjugates (Vanden Heuvel et al., 1991). The main route for elimination of PFOA is generally urinary excretion and to a small extent fecal excretion (Han et al., 2012). However, in rodents the urinary clearance rate of PFOA has been reported to be reduced due to reabsorption of the chemical from renal filtrate (Katakura et al., 2007; Lee and Schultz, 2010). Thus, lack of compound metabolism, enterohepatic recirculation and renal reabsorption are important factors that partly can explain the relatively long half-life and accumulation of PFOA. Apart from fecal and urinary clearance, another elimination pathway of chemicals is through maternal transfer into the eggs. In the present study, 14 C-PFOA labeling was observed throughout mature vitellogenic oocytes but not in immature non-vitellogenic oocytes other than labeling of the eggshell. It is known that PFOA and other PFAAs have a high affinity to blood proteins, such as albumin and liver fatty acid binding protein (Luebker et al., 2002; Scheng et al., 2014). The results of the present study could indicate that PFOA bind to major constituent yolk and eggshell proteins, such as vitellogenin and zona radiata proteins, which are synthesized in the liver and transported by the blood stream to the ovary for uptake into the maturing oocytes. This maternal transfer pathway implies adverse effects on early embryonic developmental and offspring health. In fact, in long-term studies with zebrafish exposed to PFOS adverse effects on reproduction, embryonic growth and offspring development has been reported (Wang et al., 2011). The decreased larval survival was correlated to the PFOS concentrations measured in offspring, most likely being maternally derived from the exposed parental female zebrafish (Wang et al., 2011). Maternal transfer of PFAAs has also been indicated in wild fish due to the presence of high concentrations in fish eggs (Kannan et al., 2005). Comparing toxic effect levels reported in the literature with measured levels in natural waters do not indicate a risk for acute toxicity to aquatic organisms. The concentrations of PFAAs in surface waters have usually been detected at ng/L levels (Clara et al., 2009). We and others have previously reported that the embryo toxicity of different PFAAs is generally low to zebrafish with effect concentrations commonly observed at mg/L levels (Hagenaars et al., 2011; Zheng et al., 2012; Ulhaq et al., 2013a). However, we have also observed that early embryonic exposure to PFAAs at lower concentrations can cause disturbances in behavior in zebrafish larvae (Ulhaq et al., 2013b). These findings, together with indications of accumulation of PFOA in the brain tissues can indicate a risk for PFOA to act as a potential developmental neurotoxicant. The results

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from the present study suggest that focus should be on chronic toxicity testing and long-term effects. Also, from the present study the relationship between external exposure concentrations to total body and tissue burden can be used to improve risk assessment of PFOA in the aquatic environment.

Acknowledgement The chemical synthesis of 14 C-PFOA was financially supported by unrestricted grants from the 3M Company and through faculty funding from Stockholm University.

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Tissue uptake, distribution and elimination of (14)C-PFOA in zebrafish (Danio rerio).

Perfluorooctanoic acid (PFOA) is a long-chain perfluorinated chemical that has been shown to be non-degradable and persistent in the environment. Labo...
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