Comparative Tissue Distribution and Depuration Characteristics of Copper Nanoparticles and Soluble Copper in Rainbow Trout (Oncorhynchus mykiss)

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1 Environmental Toxicology and Chemistry Volume 38, Number 1 pp , Received: 8 May 2018 Revised: 14 June 2018 Accepted: 24 September 2018 Environmental Toxicology Comparative Tissue Distribution and Depuration Characteristics of Copper Nanoparticles and Soluble Copper in Rainbow Trout (Oncorhynchus mykiss) Stina Lindh, a, * Parastoo Razmara, b Sarah Bogart, b and Gregory Pyle b a Faculty of Engineering, Lund University, Lund, Sweden b Department of Biological Sciences, University of Lethbridge, Lethbridge, Alberta, Canada Abstract: The continuously growing number of products containing nanoparticles (NPs) makes their presence in the environment inevitable, and given the well-known toxicity of dissolved metals, concerns regarding the toxicity of metal-based NPs have been raised. Whether metal-based NPs present similar or different toxicological effects compared with metal salts is an emerging field. In the present study, rainbow trout were intraperitoneally injected with CuSO 4 or copper NPs (CuNPs) to investigate tissue distribution and depuration characteristics. Fish injected with Cu showed an initial accumulation of Cu in the liver, kidney, gills, intestine, and carcass. The Cu concentration in the liver of CuNP-injected fish increased over time. It appears as though CuNPs accumulated in the liver at a greater rate than they were excreted. In livers of fish injected with CuSO 4, the Cu concentration appeared to increase and reach an equilibrium, suggesting that copper was accumulated and excreted at the same rate. The possibility that CuNPs can accumulate at a higher rate than it is excreted in the liver warrants further investigation. The present study demonstrates the possibility of dietary uptake of CuNPs because elevated Cu concentrations were observed in carcass, gills, kidney, and intestine of fish gavaged with CuNPs. In conclusion, the results suggest that dietary CuNPs are taken up by the gut and preferentially accumulate in the liver. Environ Toxicol Chem 2019;38: C 2018 SETAC. Keywords: Fish; Copper; Nanoparticles; Metal accumulation; Depuration; Tissue distribution INTRODUCTION The growing number of products containing nanoparticles (NPs) makes their presence in the environment inevitable. Household products, textiles, microelectronics, and medical appliances are examples of applications in which copper NPs (CuNPs) can be found (Mudunkotuwa and Grassian 2011). Soil leaching, atmospheric deposition, and effluent discharges are all possible routes of contamination (Scown et al. 2010). In fact, the increased usage of metallic NPs in commercial products and their subsequent release into aquatic environments have most likely resulted in growing environmental concentrations of metal-based NPs. Copper, an essential element, can be toxic to aquatic organisms at elevated concentrations. Copper sulfate is regularly used in aquaculture to control algal blooms and to inhibit ectoparasitic and bacterial infections (Griffin and Mitchell 2007). This article includes online-only Supplemental Data. * Address correspondence to atn10sli@student.lu.se Published online 1 October 2018 in Wiley Online Library (wileyonlinelibrary.com). DOI: /etc.4282 However, soluble Cu has been shown to be highly toxic to fish at elevated concentrations (Howarth and Sprague 1978). Soluble Cu can be taken up by the gills when present in the water or by the gut during dietary exposure. The mechanism of Cu uptake across the epithelium mainly involves carrier-mediated transport on metal ion transporters (Bury et al. 2003; Shaw and Handy 2011). Research into the toxicity of CuNPs is beginning to increase, especially for fish and aquatic invertebrates. When added to water, CuNPs do not exist in an aqueous solution but rather as a colloidal suspension and can aggregate, sediment out of the water column, adsorb to nutrients, and release ions, all of which may affect their bioavailability in comparison with soluble Cu (Handy et al. 2008b). Further research is needed to determine routes of CuNP uptake in fish. However, possible routes of uptake may be direct passage through the gills or through the epithelial cells of the intestine. As opposed to Cu ions, CuNPs are too large to be taken up by ion, protein, or transepithelial transporters and possibly endocytosis (Shaw and Handy 2011). A feature that has long been exploited in the oral delivery of fish vaccines, is the ability of the intestinal epithelium in fish to take up much larger materials by endocytosis compared with the intestinal epithelium of mammals (Handy et al. 2008a; Løvmo et al. 2017). C 2018 SETAC wileyonlinelibrary.com/etc

2 Comparative tissue distribution of Cu and CuNP in trout Environmental Toxicology and Chemistry, 2019;38: Consequently, fish may be more vulnerable to dietary CuNP exposure relative to mammals. Knowledge of CuNP uptake routes is extremely limited compared with those of soluble Cu. Several studies focusing on the toxicity of soluble Cu to fish exposed by various routes, such as water or diet, exist (Miller et al. 1993; Kamunde et al. 2002). Studies have shown that soluble Cu can be effectively taken up by the gills and the intestine and that soluble Cu is toxic to fish at elevated concentrations (Miller et al. 1993; Kamunde et al. 2002). Further, fish exposed to elevated concentrations of soluble Cu can accumulate a high proportion of the newly acquired Cu in the liver, gills, and kidney (Grosell et al. 2001). Currently, little is known about the toxicity of CuNPs, and research is needed to determine uptake routes, tissue distribution, and depuration characteristics of CuNPs. In the present study, the term depuration includes the initial period of uptake and distribution to tissues as well as the depuration of Cu/CuNPs. A few studies have demonstrated the possibility of CuNP uptake through the gills, but the mechanism is unknown (Shaw et al. 2012). To the best of our knowledge, the possibility of CuNP uptake through the intestine and the tissue distribution and depuration characteristics of CuNPs have yet to be determined. The main aim of the present study was to compare tissue distribution and depuration characteristics of soluble Cu and CuNPs in rainbow trout. Because several factors may affect Cu uptake by natural routes, rainbow trout were injected intraperitoneally with nonlethal doses of either CuSO 4 or CuNPs. The present study also evaluated natural routes for potential uptake of CuNPs, by exposing fish in a waterborne regime or gavaging them with CuNPs. FIGURE 1: Particle size and particle size distribution of 10 mg/l copper nanoparticle (CuNP) stock. (A) Electron micrograph showing aggregated CuNPs with a mean particle size of nm (mean standard error of the mean, n ¼ 200). (B) Size distribution of CuNPs. MATERIALS AND METHODS CuNP characterization The CuNPs used were purchased from NanoAmor and had an average particle size of 35 nm and 99.8% purity. The size and shape of the CuNPs were confirmed using transmission electron microscopy (TEM; FEI Tecnai-20; Figure 1A) in 10 mg L 1 dilutions (in Millipore water). The corresponding size distribution of CuNPs (10 mg L 1 ) was measured using the image analyzing and processing software ImageJ (Dorobantu et al. 2015) based on TEM images (Figure 1B). The primary average diameter of CuNPs was nm (mean [standard error of the mean] SEM, n ¼ 200). To determine the aggregation of CuNPs, experimental suspensions (45, 120, and 320 mgl 1 ) over time (24 h), dynamic light scattering (DLS; Zetasizer Nano Series; Malvern) was used in incipient light angle backscatter mode (n ¼ 3; Zetasizer software, Ver 7.01). The average hydrodynamic diameter, polydispersity, and zeta potential of CuNPs were measured at 0, 2, 4, 8, 16, and 24 h (Table 1). The CuNP suspensions were prepared using fish tank water (ph , conductivity ms cm 1, temperature C, dissolved oxygen mg L 1, dissolved organic carbon mg L 1, and total hardness 158 2mgL 1 as CaCO 3 ). Between the DLS measurements, samples were covered by Parafilm and remained static. The results of light scattering measurements indicated that CuNPs tend to aggregate over time. However, there were no significant differences in respect to aggregation behavior of CuNPs among different exposure concentrations over time (Table 1; p > 0.05). The polydispersity index (PDI), an indicator of particles relative dispersity in the suspension, was significantly increased over time, which demonstrates that particles were highly polydisperse and aggregated over time (Table 1; p 0.001). Like the PDI, the hydrodynamic diameter was significantly increased over time and confirmed the aggregation behavior of CuNPs over 24 h (Table 1; p ¼ 0.041). The zeta potential of CuNPs was not affected over time (p > 0.05). The dissolution profile of CuNPs at 0.25 g L 1 30 min after preparing the stock suspension was investigated. A stock suspension was prepared in water at a concentration of 0.25 mgl 1 and sonicated for 30 min. Thirty minutes after sonication, samples were centrifuged at rpm for 30 min at 4 8C, and subsequently the supernatants were filtered through a mm syringe filter (Whatman; catalog no ; Xiao et al. 2015). The percentage of ions released from the stock solution was %. wileyonlinelibrary.com/etc C 2018 SETAC

3 82 Environmental Toxicology and Chemistry, 2019;38:80 89 S. Lindh et al. TABLE 1: Copper nanoparticle characterizations: polydispersity index, hydrodynamic diameter, and zeta potential at 3 nominal concentrations (45, 120, and 320 mg L 1 ) measured by dynamic light scattering at 0, 2, 4, 8, 16, and 24 h (mean standard error of the mean, n ¼ 3) Time (h) Nominal concentration (mgl 1 ) PDI HDD (nm) Zeta potential (mv) Significant increase of polydispersity index or hydrodynamic diameter over time. Preparation of CuNP stock suspension and CuSO 4 stock solution For fish exposure experiments, CuNP stock suspensions and CuSO 4 stock solutions were prepared as follows. Intraperitoneal injection. A fresh stock was prepared prior to each injection by suspending or dissolving CuNP or CuSO 4, respectively, in physiological saline at a Cu concentration of 0.25 g L 1. Each fish was injected with 10 ml/g of fish, resulting in an injection concentration of 2.5 mg Cu/g of fish. To avoid aggregation of the CuNPs, suspensions were sonicated for 30 min prior to intraperitoneal injection. The CuSO 4 solution and the sham (physiological saline) were also sonicated for 30 min prior to injection, to make sure that all of the solutions used for injection were prepared in the same way. Waterborne exposure. A fresh 0.5 g L 1 CuNP stock suspension was prepared daily prior to water change in Millipore water. To avoid aggregation, the suspension was sonicated for 30 min prior to dosing of experimental aquariums. Preparation of food paste for gastric gavage A food paste was prepared by soaking commercial trout chow in Millipore water (trout chow: Millipore water 1:1 w/v) overnight. On the day of the gavage, the soaked trout chow was thoroughly mixed, using glass rods as pestles and 50 ml tracemetal-free, polypropylene digestion tubes (Agilent Technologies) as mortars, and Cu was added as either CuSO 4 or CuNPs or not at all and thoroughly mixed using glass rods. Fish husbandry Three sets of experiments were performed: an intraperitoneal injection of either CuNPs or CuSO 4 to investigate and compare tissue distribution and depuration characteristics of Cu for both CuNPs and Cu ions, a waterborne exposure to determine the uptake of CuNPs by the gills, and a gastric gavage to evaluate the potential uptake of CuNP through the gastrointestinal tract. Juvenile rainbow trout (n ¼ 323), weighing g (mean SEM, n ¼ 163) were obtained from Sam Livingstone Fish Hatchery and housed in a holding tank containing 12 8C dechlorinated facility water under a 16:8-h light:dark photoperiod for a minimum of 2 wk prior to experimentation. All procedures in the present study involving fish were approved by the University of Lethbridge Animal Welfare Committee under the guidelines of the Canadian Council for Animal Care. Experimental design Intraperitoneal injection. At experiment day 0, 48 rainbow trout were anesthetized using a ph-buffered MS-222 solution (200 mg L 1 MS-222; 600 mg L 1 NaHCO 3 ) and intraperitoneally injected or not with either a control (no injection), a sham (injected with physiological saline), or 2.5 mg Cu/g body weight, as either CuSO 4 or CuNPs (n ¼ 12 fish/treatment). The injection volume was 10 ml/g body weight. Following intraperitoneal injection, fish were transferred into 16 aerated experimental glass aquariums (3 fish/tank) containing 20 L dechlorinated facility water. At 24 h (n ¼ 3 fish for each treatment), 48 h (n ¼ 3 fish for each treatment), 96 h (n ¼ 3 fish for each treatment), and 192 h (n ¼ 3 fish for each treatment) postinjection, fish were euthanized; and tissues (liver, kidney, gills, intestine, and carcass) were immediately harvested and placed in glass vials (liver, kidney, and gills) or glass Petri dishes (carcass). The intestine was rinsed in physiological saline and subsequently placed in a glass vial. Samples were oven-dried to constant mass at 60 8C. The experiment was repeated to obtain n ¼ 6 fish/treatment/time point (96 fish in total). Waterborne exposure. Sixteen rainbow trout were transferred into 8 aerated experimental glass aquariums (2 fish/tank) containing 20 L dechlorinated facility water, in a duplicate design C 2018 SETAC wileyonlinelibrary.com/etc

4 Comparative tissue distribution of Cu and CuNP in trout Environmental Toxicology and Chemistry, 2019;38: (2 tanks/treatment). Fish were exposed for 96 h using a semistatic exposure regime (a 60% water change was performed every 24 h with redosing following each change) to either control (no added CuNP) or Cu treatment of 45, 120, or 320 mgl 1 CuNPs. The exposure concentrations of CuNPs were chosen based on the expectation that they would result in sublethal effects. Specifically, Song et al. (2015) reported a 96-h median lethal concentration (LC50) of 0.68 mg L 1 for CuNPs in rainbow trout. To reduce the risk of having concentration-dependent mortalities, 60% of the LC50 value reported in Song et al. (2015) was used as the highest nominal test concentration (400 mgl 1 CuNPs), which resulted in a measured exposure concentration of 320 mgl 1 CuNPs. The lowest nominal test concentration (70 mgl 1 ) was chosen as 10% of the LC50 reported in Song et al. (2015), which resulted in a measured exposure concentration of 45 mgl 1 CuNPs. The remaining exposure concentrations were calculated based on a geometric dilution series. Fish were not exposed to CuSO 4 in a waterborne regime because studies have shown that soluble Cu causes mortality at concentrations much lower than the concentrations of CuNPs used in the present study (Shaw et al. 2012). Further, the objective of the present study was to evaluate if the toxicity of CuNPs could be explained by the dissolution of CuNPs into Cu ions. However, the weak dissolution profile of CuNPs and, hence, the low concentration of Cu ions rendered the realization of the present study less feasible; therefore, waterborne studies investigating the toxicity of soluble Cu were not performed. Water samples were collected at the start (0 h), before and after each water change, and at the end (96 h). Water samples collected at the start (0 h) and at the end (96 h) of the exposure were analyzed to determine exposure concentrations. Fish were euthanized at 96 h, and tissues were sampled as described for intraperitoneal injection. Gastric gavage. At experimental day 0, 12 rainbow trout were divided into 4 groups (n ¼ 3/treatment), anesthetized using a ph-buffered MS-222 solution (200 mg L 1 MS-222; 600 mg L 1 NaHCO 3 ), and gavaged or not with either a control (no perfusion), a sham (0.09 mg Cu/g body wt; no added Cu to the perfused food paste), 200 mgcu/gbodyweightofcuso 4 (measured concentration 50 mgcu/gbodywt),or200mg Cu/g body weight of CuNPs (measured concentration 120 mg Cu/g body wt). The perfusion volume was 10 ml/g body weight (according to Canadian Council on Animal Care Guidelines Committee [2005] guidelines). The gavage apparatus consisted of an 18-G plastic tube, cut to approximately 5 cm, and placed on an 18-G needle attached to a 1-mL syringe. Following the gavage, fish were placed in a recovery tank for 30 min and monitored for any signs of regurgitation. After the 30-min recovery, fish were transferred into 4 aerated experimental glass aquariums (3 fish/tank) containing 20 L dechlorinated facility water. At 24 h, fish were transferred to clean glass aquariums, and a 100% water change was performed. Fish were euthanized at 48 h, and tissues were sampled as described for intraperitoneal injection. Fish were restricted from food 24 h prior to and 24 h after the start of exposure. Starting at 48 h after the start of exposure, fish were fed commercial trout chow every second day at a basic maintenance diet of 0.5% of their body weight until the desired endpoint was reached. Water quality analysis Methods and results of the water quality analysis are given in the Supplemental Data. Sample preparation for Cu analysis Dried fish tissue samples were homogenized using a mortar and pestle (which were thoroughly rinsed with methanol between experiment duplicates and acid-washed between samples from different tissue types and treatments) and subsequently aciddigested to permit Cu analysis. Because of the small size of the rainbow trout used in the present study, tissue samples collected at the end of the study were small. Thus, microdigestions of fish tissues were completed where, depending on the amount of fish tissue available, aliquots of certified reference material (CRM; i.e., DOLT-4 dogfish liver; National Research Council Canada) were weight-matched to the fish tissue samples to evaluate the accuracy of the sample digestion method. In short, aliquots ranging from 10 to 3 mg of dried tissue were weighed into 2-mL microcentrifuge tubes with locking caps (Fisher Scientific; catalog no ). Concentrated nitric acid (TraceSELECT grade; Honeywell; catalog no ml) was added to the preweighed samples in a 1:10 ratio (tissue mass [milligrams]: acid volume [microliters]), and the lids were tightly closed. Samples were subsequently digested in a hot-block for 3 h at 80 8C (Jung 2008). Three acid-only tubes (blanks) were digested simultaneously during each digest run to confirm the Cu cleanliness of the digest procedure. A CRM was run every twenty-second sample or a minimum of 3 CRM per digest run depending on the number of samples being digested. Postdigest, solutions were allowed to cool to room temperature and subsequently made up to 2 ml volume with Millipore water (Jung 2008). All samples were then analyzed for Cu concentration (see below, Cu analysis). Mean percentage of recovery of Cu in DOLT- 4 across all microdigests was 89.9% (range %). Similarly, Cu concentrations in the food paste used for gastric gavage were determined by digesting aliquots of approximately 100 mg food paste samples (n ¼ 1 sample/treatment) and weight-matched CRM for 3 h at 80 8C with 1 ml concentrated nitric acid in glass digestion tubes. Three blanks and 3 CRMs were digested simultaneously during the run. The digested solutions were allowed to cool to room temperature and subsequently made up to 20 ml volume with Millipore water. All samples were then analyzed for Cu concentration (see below, Cu analysis). In the food digest, mean percentage of recovery of Cu in DOLT-4 was 81.7% (range %). Cu analysis All Cu concentrations were determined by an Agilent 240 FS atomic absorption spectrometer equipped with a GTA 120 graphite tube atomizer (Agilent Technologies). Samples were analyzed at a wavelength of nm and a slit width of 0.5 nm, in background correction mode, using either a Cu hollowcathode lamp or a multielement HC lamp (Agilent Technologies; catalog no or ). An extra step was wileyonlinelibrary.com/etc C 2018 SETAC

5 84 Environmental Toxicology and Chemistry, 2019;38:80 89 S. Lindh et al. added to the furnace burn-profile (step 9, temperature C, time 4.0 s, flow 0.3 L/min), to reduce carryover of previous sample; otherwise, sample analysis was according to the manufacturer. All samples were diluted as necessary to fall within the calibration range (10 90 mgl 1 Cu). The CRM SLRS-6 (National Research Council Canada) was run every 10 samples to evaluate the accuracy of the analysis (mean percentage of recovery 99.0%, range %). Analytical duplicates were run approximately every 25 samples (mean difference 0.62%). The detection limit for Cu was estimated during sample analysis to be approximately 2 mgl 1. Statistical analysis All data were analyzed using R, Ver (R Development Core Team 2016). Data obtained from the intraperitoneal injection were analyzed for main effects of treatment, time, and treatment time interaction effects by permutation multivariate analysis of variance (MANOVA). Effects that were statistically significant were analyzed by a one-way permutation analysis of variance (ANOVA). Tukey s post hoc tests were performed to observe the effects of treatment at each time point, whereas Dunnett s post hoc tests were performed to analyze how the Cu concentrations at 48, 96, and 192 h varied compared to the Cu concentrations at 24 h for each of the 4 treatments. Data obtained from the waterborne exposure were analyzed using MANOVA. Effects that were statistically significant were analyzed by a one-way ANOVA followed by a Dunnett s post hoc test. Data obtained from the gastric gavage were analyzed using permutation MANOVA. Effects that were statistically significant were analyzed by a one-way permutation ANOVA followed by a Dunnett s post hoc test. Differences among groups were considered significant when p Grubbs test was performed on the liver data obtained for the intraperitoneal injection to test for the possibility of outliers. One liver sample contained an extremely high concentration of Cu, was considered an outlier at p 0.05, and was removed before further statistical analysis. RESULTS Intraperitoneal injection Copper concentrations of water samples taken at the start (0 h) and at the end (192 h) of the intraperitoneal injection experiment and stock solutions and suspensions were confirmed by graphite furnace atomic absorption spectrophotometry (GFAAS; Table 2). The variation in Cu stock concentration (within-treatment duplicates) for the intraperitoneal injection was <3%, suggesting that the concentrations of exposure solutions were maintained throughout the experiment. Data from treatment duplicates were pooled, giving the Cu concentrations shown in Table 2. Copper concentrations of water samples collected at the start (0 h) and at the end (192 h) of the experiment were all below the detection limit of the instrument. Fish injected with either Cu as CuSO 4 or CuNPs showed an accumulation of Cu in all tissues relative to controls (p ¼ 0.001; Figure 2), with the highest increase of Cu observed in the kidney at 30 and 40 times more Cu compared to controls for CuSO 4 - and CuNP-injected fish, respectively. Copper concentrations in the kidney of Cu-injected fish did not change over time (permutation ANOVA; CuSO 4, p ¼ 0.81; CuNP, p ¼ 0.15; Figure 2C). The highest concentration of Cu was measured in the liver of Cu-injected fish, and concentrations were approximately doubled relative to controls for both treatments. An interesting observation was that CuNP-injected fish showed a statistically significant increase of Cu concentration in the liver at 192 h compared to 24 h postinjection (permutation ANOVA, p < ; Figure 2B). A significant accumulation of Cu over time (permutation ANOVA, p < 0.05; Figure 2B) was also observed for CuSO 4 -injected fish. However, the accumulation of Cu in the liver of fish injected with CuSO 4 seemed to increase at first before reaching an equilibrium. The carcass, gills, and intestine all showed a reduction of Cu over time (permutation ANOVA, p < 0.05; Figure 2A,D,E), with Cu concentrations in the gills almost equivalent to the controls at 192 h. Grubbs test was performed on the data obtained from the liver to detect possible outliers. The liver sample showing the highest measured Cu concentration was an outlier according to the test (p < ). The measured concentration was approximately 4 times higher in one of the liver samples compared to the replicates of the same treatment. The sample containing the outlier influenced the mean Cu concentration substantially. Consequently, the outlier was removed prior to statistical analysis. Waterborne exposure Waterborne copper exposure concentrations were confirmed by GFAAS analysis of water samples taken at the start TABLE 2: Nominal and measured copper concentrations of the stock solutions used for the intraperitoneal injection a Exposure Nominal Cu concentration (g L 1 ) Measured average Cu concentration (g L 1 ) Percentage of nominal Sham Below detection limit Mean Below detection limit NA CuNP 0.25 Mean SEM Cu as CuSO Mean SEM a Nominal and measured average copper concentrations with standard error of the mean of the sham (n ¼ 2), CuNP (n ¼ 8), and CuSO 4 (n ¼ 4) stock solutions used for the intraperitoneal injection. Also included are the percentage of the nominal Cu concentration for each treatment. CuNP ¼ copper nanoparticle. C 2018 SETAC wileyonlinelibrary.com/etc

6 Comparative tissue distribution of Cu and CuNP in trout Environmental Toxicology and Chemistry, 2019;38: FIGURE 2: Copper concentrations in (A) carcass, (B) liver, (C) kidney, (D) gills, and (E) intestine of trout 24, 48, 96, and 192 h after intraperitoneal injection with either physiological saline (sham), Cu as CuSO 4, copper nanoparticles, or no injection (control). Data are mean standard error of the mean, dry weight tissue, n ¼ 5 to 6 fish/treatment and time point. Different letter denotes a statistically significant difference between treatments at each time point (p < 0.05); significantly different from same treatment at the time point 24 h (p < 0.05). Grubbs test was performed on the data obtained for the liver to detect possible outliers. The highest measured concentration was according to the test an outlier (p < 0.001) and was therefore removed before any further statistical analyses were performed. CuNP ¼ copper nanoparticle. (0 h) and at the end (96 h) of the experiment. The variation in Cu concentration within treatment duplicates at 0 and 96 h and the variation between 0 and 96 h were <5%, confirming that exposure concentrations were maintained throughout the experiment. Data from treatment duplicates were pooled, giving the Cu concentrations shown in Table 3. Copper concentrations for the control were below the detection limit of the instrument. Copper concentrations were marginally elevated in some fish tissues relative to others after a 96-h exposure to CuNP (p ¼ 0.08). In particular, gill Cu increased by more than 2-fold at all exposure concentrations tested (ANOVA; p ¼ 0.003; Figure 3D) relative to controls. Intestinal Cu concentrations were also significantly elevated but only in fish exposed to the highest CuNP concentration (ANOVA; intestine, p ¼ 0.047; Figure 3E). No other tissues accumulated significantly more Cu in fish exposed to CuNP than those in controls (ANOVA; p > 0.05; Figure 3A C). Gastric gavage Copper concentrations in the water at 0, 24, and 48 h were determined by GFAAS analysis of water samples. The Cu concentrations in water from control, sham, and water samples taken at 0 h were below the detection limit of the instrument. Analysis of water samples taken in glass aquariums holding fish gavaged with CuNPs revealed Cu concentrations of and 5.2 mgl 1 at 24 and 48 h. The Cu concentration in water samples taken from glass aquariums holding fish gavaged with CuSO 4 were 26.8 and 10.3 mgl 1 at 24 and 48 h. Copper concentrations TABLE 3: Nominal and measured copper concentrations of the water samples for the waterborne exposure a Nominal Cu concentration (mgl 1 CuNP) Measured Cu concentration (mgl 1 CuNP) Percentage of nominal relative to measured average Cu concentration Control Below detection limit a Nominal and measured copper concentrations of the water samples collected in the waterborne exposure (n ¼ 4). Also included are percent of the nominal Cu concentration relative to measured average Cu concentration for each exposure. CuNP ¼ copper nanoparticle. wileyonlinelibrary.com/etc C 2018 SETAC

7 86 Environmental Toxicology and Chemistry, 2019;38:80 89 S. Lindh et al. FIGURE 3: Copper concentrations in (A) carcass, (B) liver, (C) kidney, (D) gills, and (E) intestine of rainbow trout after a 96-h waterborne exposure to one of 3 concentrations of copper nanoparticles or control. Data are mean standard error of the mean (SEM), dry weight tissue, n ¼ 3 to 4 fish/treatment. Mean and SEM could not be calculated for the liver (B) of fish exposed to 320 mgl 1 as a result of sample loss during the acid digestion of the samples (n ¼ 2). Result of multivariate analysis of variance, p ¼ 0.088; statistical significance compared to control (p < 0.05). of the food paste were determined by GFAAS analysis of food paste digest samples, and the results are given in Table 4. Fish gavaged with Cu showed an initial accumulation of Cu in the carcass, kidney, gills, and intestine relative to controls (ANOVA; carcass, p ¼ 0.01; kidney, p ¼ 0.042; gills, p ¼ 0.001; intestine, p ¼ 0.004; Figure 4). No Cu accumulation was observed for the liver (ANOVA, p ¼ 0.19; Figure 4), but a trend toward elevated concentrations was indicated by the results. DISCUSSION AND CONCLUSION The present study is, to our knowledge, one of the first reports detailing the tissue distribution and depuration characteristics of Cu in rainbow trout intraperitoneally injected with CuNPs compared to an equal mass of Cu injected as CuSO 4. The present study is, to our knowledge, also one of the first reports demonstrating the possibility of dietary uptake of CuNPs in rainbow trout. The potential that CuNPs may result in different toxicological effects compared to Cu salts makes it important to better understand the toxicology of NPs. Several factors have an impact on Cu uptake by natural routes. Therefore, to study the tissue distribution and depuration of both Cu ions and CuNPs in fish, rainbow trout were injected intraperitoneally with nonlethal doses of either CuSO 4 or CuNPs. Our results showed that Cu accumulated in the carcass, liver, kidney, gills, and intestine of fish injected with Cu and that there was no difference in the accumulation pattern between the fish injected with either CuSO 4 or CuNPs (Figure 2). The present results are in contrast to previous work, which has shown that Cu injected as CuSO 4 accumulated to higher concentrations in some tissues than copper oxide (CuO) NPs (Isani et al. 2013). The difference between the 2 studies suggests that the chemical characteristics of the NPs (size and surface coating) and/or the amount of Cu injected may influence the fate of the NPs inside the fish. The trends for carcass, gills, and intestine were a reduction in Cu concentration over 192 h for fish injected with either CuSO 4 or CuNPs. The results suggest that the Cu is redistributed to other tissues, such as the liver where an increase of CuNPs was observed over the 192 h or, in the case of the gills, possibly excreted (Figure 2A,D,E). The gills have been shown to play a key role in Cu TABLE 4: Nominal and measured copper concentrations of the food paste used for gastric gavage a Exposure Nominal Cu concentration (g L 1 ) Measured Cu concentration (g L 1 ) Percentage of nominal Sham NA 0.09 NA CuNP Cu as CuSO a Nominal and measured copper concentrations of the food paste used to gavage fish (n ¼ 1). Also included are percentage of the nominal Cu concentration for each treatment. CuNP ¼ copper nanoparticle. C 2018 SETAC wileyonlinelibrary.com/etc

8 Comparative tissue distribution of Cu and CuNP in trout Environmental Toxicology and Chemistry, 2019;38: FIGURE 4: Copper concentrations in (A) carcass, (B) liver, (C) kidney, (D) gills, and (E) intestine of rainbow trout gavaged with copper nanoparticles, CuSO 4, sham, or control (no perfusion). Data are mean standard error of the mean (SEM), dry weight tissue, n ¼ 3 fish/treatment. Mean and SEM could not be calculated for the gills (D) of fish gavaged with CuSO 4 as a result of sample loss during the acid digestion of the samples (n ¼ 2). Statistical significance compared to control (p < 0.05). homeostasis, and previous reports support the possibility of branchial Cu excretion (Grosell et al. 2001; Kamunde et al. 2002). However, caution needs to be taken when drawing a conclusion for CuNPs because the results demonstrated by Grosell et al. (2001) and Kamunde et al. (2002) are from fish exposed to soluble Cu. Fish are not known to be able to perform extensive vesicular trafficking from the basolateral to the apical surface of the gill epithelial cells (Handy et al. 2008a). Excretion of CuNP from the gills therefore seems unlikely. In fish, the liver is the main homeostatic organ for Cu metabolism. The liver has been shown to accumulate a large proportion of newly accumulated Cu, implying its important role in Cu homeostasis (Miller et al. 1993; Grosell et al. 1998, 2001). Further, high concentrations of Cu in the bile of Cu-exposed fish imply the importance of hepatobiliary excretion of excess Cu in fish (Grosell et al. 2001). The mechanism of Cu excretion by the liver relies on vesicular trafficking to form the bile. Previous work has shown that vesicles may be up to 200 nm in diameter, rendering CuNP excretion from the liver at least feasible (Hampton et al. 1988). In accordance with previous work, the present study showed that a large amount of the injected Cu is accumulated in the liver (Figure 2B). Further, liver Cu concentrations increased over the 192-h postinjection period, suggesting that excess Cu may be mobilized to the liver for excretion (Figure 2B). The increase in Cu concentration over the 192-h postinjection period was more pronounced in fish injected with CuNPs compared to fish injected with CuSO 4. The slower kinetics of Cu clearance in the liver may be a result of excess Cu being redistributed to the liver from other organs, owing to the detoxification and storage processes typical of the liver. It is possible that when a normal Cu level is maintained in other tissues, the redistribution of Cu to the liver will level off and excess Cu in the liver will be excreted, resulting in a reduction of Cu concentrations in the liver. Extending the postinjection monitoring period for the study will be desirable to define the depuration phase of both CuNPs and CuSO 4 in the liver. To the best of our knowledge, the ability of fish to clear CuNPs has yet to be determined. In general, bioaccumulation is not considered a major factor for soluble Cu (Suedel et al. 1994), presumably owing to the relatively strong homeostatic control of this essential element. However, the data obtained in the present study cannot exclude the possibility of bioaccumulation being an important factor for CuNPs. It is likely thattheobservedincreaseofcuintheliverofbothcunp-and CuSO 4 -injected fish is a result of excess Cu being redistributed to the liver. However, the ability to excrete CuNPs and CuSO 4 by the liver seems to differ. A statistically significant increase of Cu over time was observed in the liver of both CuNP- and CuSO 4 -injected fish. However, in the liver of CuSO 4 -injected fish the accumulation of Cu seemed to increase at the start and then reach an equilibrium, whereas for CuNP-injected fish the accumulation did not seem to level off and reach an equilibrium over the 192 h, rendering bioaccumulation of CuNPs a possible outcome. Because intraperitoneal injection is different from natural contamination, it may not necessarily stimulate the same biological responses as a direct exposure to Cu in the field. Bioaccumulative properties are highly dependent on the ability of fish to take up CuNPs from the environment and can therefore not be determined by intraperitoneal injection. However, the present data preliminarily indicate that CuNPs may have the potential to bioaccumulate, and this should wileyonlinelibrary.com/etc C 2018 SETAC

9 88 Environmental Toxicology and Chemistry, 2019;38:80 89 S. Lindh et al. be investigated further. During dissection of the fish it was observed that the bile from fish injected with Cu as CuSO 4 was much darker than the bile from fish injected with CuNPs, control, or sham, suggesting that the contents of the bile may differ among treatments. Further research, with an extended postinjection monitoring period and investigation of the Cu concentration in the bile, may allow for a better understanding of the bioaccumulative properties of CuNPs. The kidney was also a target, with elevated Cu concentrations in fish injected with Cu; however, the length of the time period from intraperitoneal injection to sampling had no effect on kidney Cu concentration (Figure 2C). In freshwater teleosts, urinary Cu excretion is extremely low, probably as a result of the tight association between Cu and plasma protein, suggesting that Cu is less available for glomerular filtration (Grosell et al. 1998; Grosell 2012). A potential explanation for the slower kinetics of Cu clearance in the kidney may be the reabsorption of hepatic Cu-metallothionein followed by renal thionein synthesis (Isani et al. 2013). The waterborne exposure showed a significant increase of Cu concentrations in the gills of fish exposed to CuNPs for 96 h, suggesting that CuNPs may be taken up by the gills (Figure 3D). The results are in agreement with those of Griffitth et al. (2007) and Shaw et al. (2012), who demonstrated Cu accumulation in the gills of zebrafish and rainbow trout, respectively, after exposure to 100 mgl 1 CuNPs. No observable Cu accumulation in the liver of CuNP-exposed fish suggests that the gills are not effectively taking up CuNPs from the surrounding environment because newly acquired Cu has been shown to accumulate quickly in the liver (not observed in the present study; Grosell et al. 2001). The possibility that CuNPs are not taken up effectively by the gills is further supported by the lack of concentration response in the kidney (Figure 3C). No observed concentration response in internal organs supports the possibility that CuNPs are not internalized. Association of CuNPs with the mucus in the gill microenvironment (seen for other NPs; Smith et al. 2007), as a result of steric hindrance or binding to mucus proteins, and the suggested mechanism of uptake through endocytosis may partly explain the low accumulation of CuNPs in internal organs (Handy et al. 2008a; Shaw and Handy 2011). In addition, association of CuNP with the mucus renders it possible that the CuNPs are not taken up by the gills but rather associated with the mucus layer. The possibility that CuNPs are associated with the mucus rather than taken up by the gills is further supported by the lack of concentration response. It is possible that the mucus is saturated with CuNPs even at low concentrations and, hence, no concentration response will be observed. Of the other tissues analyzed for Cu, only the intestine of fish exposed to 320 mgl 1 CuNPs displayed significantly elevated Cu concentrations (Figure 3E). A possible explanation for the increased accumulation of Cu in the intestine may be ingestion of CuNPs, which possibly had adsorbed to the surface of the food pellets and, hence, ingested while the fish were feeding. The elevated Cu concentrations in glass aquariums holding fish gavaged with Cu render the occurrence of regurgitation after the transfer of fish from recovery tanks to experimental glass aquariums a possible outcome. However, the elevated Cu concentrations at 48 h (after transfer to a clean aquarium and a 100% water change at 24 h) suggest that at least a portion of the gavaged Cu was maintained within the fish because regurgitation after 24 h is most unlikely. The possibility that parts of the gavaged food paste were regurgitated and the fact that fish, in addition to a possible dietary exposure, were exposed through a waterborne regime render the drawing of conclusions regarding gastrointestinal uptake difficult because it cannot be determined whether the accumulated Cu is a result of a dietary exposure, a waterborne exposure, or both. The measured concentration of Cu in the food paste (50 and 120 mg Cu/g body wt for CuSO 4 and CuNPs, respectively) differed from the nominal Cu concentration (200 mg Cu/g body wt for both CuSO 4 and CuNPs). The error may be a result of insufficient digestion of the food sample or of the amount of Cu added to the food paste during preparation being too low. The low percentage of nominal concentration in this experiment is an experimental weakness and will be corrected in future studies. The data are being reported because they suggest that dietary uptake of CuNPs is a possible uptake route for CuNPs. However, to draw conclusions regarding dietary uptake of CuNPs, additional studies are required. Both branchial uptake and intestinal uptake have been demonstrated for CuSO 4 ; hence, the accumulated Cu in fish gavaged with CuSO 4 is likely a result of both dietary and waterborne exposures (Kamunde et al. 2002; Kamunde and MacPhail 2008). The present study showed that CuNPs were significantly accumulated in only the gills and intestine after exposure to 320 mgl 1 CuNPs for 96 h in a waterborne regime (Figure 3). The present study also showed that CuNPs accumulate in the carcass, kidney, gills, and intestine, with a trend toward elevated Cu concentrations in the liver of fish gavaged with CuNPs (Figure 4). Because water samples from fish gavaged with CuNPs showed a highest measured CuNP concentration of mgl 1 and exposure to 320 mgl 1 CuNPs for 96 h in a waterborne regime only resulted in elevated tissue Cu concentrations in the gills and intestine, the results suggest that at least a fraction of the gavaged food paste remained within the fish. This result further suggests that direct passage through the epithelial cells of the intestine may be a possible route of CuNP uptake. A trait that has long been exploited in the oral delivery of fish vaccines is the ability of the intestinal epithelium in fish to take up much larger materials by endocytosis compared with the intestinal epithelium of mammals. This trait may render fish more vulnerable to dietary exposure than mammals (Handy et al. 2008a; Løvmo et al. 2017). It is likely that the binding of CuNPs to the mucus contributed to the observed accumulation of CuNPs in the intestine. However, the observed elevated Cu concentrations in carcass, kidney, and gills and the observed trend toward elevated Cu concentrations in the liver suggest that the accumulation of CuNPs in the intestine cannot solely be explained by CuNPs binding to the mucus. The observed elevated Cu concentration in the intestine of fish gavaged with CuNPs is likely a combination of CuNPs associating with the mucus and intestinal uptake of CuNPs. It therefore seems possible that dietary exposure to CuNPs by rainbow trout may result in gastrointestinal uptake of CuNPs, leading to accumulation of CuNPs in organs and tissues. However, further studies are needed to draw conclusions C 2018 SETAC wileyonlinelibrary.com/etc

10 Comparative tissue distribution of Cu and CuNP in trout Environmental Toxicology and Chemistry, 2019;38: regarding uptake, tissue distribution, toxicity, and depuration of dietary CuNPs. In conclusion, Cu, whether in the form of an ion or a NP, was, after intraperitoneal injection, accumulated in the rainbow trout carcass, liver, kidney, gills, and intestine with no observable difference in Cu concentrations between the 2 treatments. However, the liver s ability to excrete CuNPs and CuSO 4 seems to differ. A statistically significant increase of Cu over time was observed in the liver of both CuNP- and CuSO 4 -injected fish. However, in the liver of CuSO 4 -injected fish the accumulation of Cu seemed to increase at the start and then reach an equilibrium, whereas for CuNP-injected fish the accumulation did not seem to level off and reach an equilibrium over the 192 h, rendering bioaccumulation of CuNPs a possible outcome. Thus, the possibility of CuNP bioaccumulation is a field deserving attention in future studies. Further, the uptake of CuNPs in a waterborne regime appears to be limited because an elevated Cu concentration in fish was observed almost exclusively in the gills after waterborne CuNP exposure. However, dietary uptake of CuNPs appears to be a possible route. The likely occurrence of CuNPs adsorbing to food may increase the possibility of dietary uptake in fish, thereby potentially rendering fish vulnerable to exposure to CuNPs both in a dietary exposure regime, as in the present study, and in a chronic waterborne exposure during which fish are being fed. The possibility and extent of CuNP uptake by the gut, as well as the potential for toxicological effects, is a field deserving further attention. Supplemental Data The Supplemental Data are available on the Wiley Online Library at DOI: /etc Acknowledgment We thank E. Lari and E. Stock at the University of Lethbridge for their kind support throughout the present study. This research was partially funded by a Natural Sciences and Engineering Research Council (NSERC) Discovery Grant to G.G. Pyle (grant number STPGP ). G.G. Pyle is also supported by a Campus Alberta Innovation Program (CAIP) chair in Aquatic Health. Data Accessibility Data pertaining to the present study are available in a publicly available data repository. To get access, please contact Dr. Gregory Pyle at gregory.pyle@uleth.ca. REFERENCES Bury NR, Walker PA, Glover CN Nutritive metal uptake in teleost fish. J Exp Biol 206(Pt. 1): Canadian Council on Animal Care Guidelines Committee CCAC guidelines on the care and use of fish in research, teaching and testing. Canadian Council on Animal Care, Ottawa, ON, Canada. Dorobantu LS, Fallone C, Noble AJ, Veinot J, Ma G, Goss GG, Burrell RE Toxicity of silver nanoparticles against bacteria, yeast, and algae. Journal of Nanoparticle Research 17:172. 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