Project description:Ketoprofen and diclofenac are non-steroidal anti-inflammatory drugs (NSAIDs) often used for similar indications, and both are frequently found in surface waters. Diclofenac affects organ histology and gene expression in fish at around 1 M-BM-5g/L. Here, we exposed rainbow trout to ketoprofen (1, 10 and 100 M-BM-5g/L) to investigate if this alternative causes less risk for pharmacological responses in fish. The bioconcentration factor from water to fish blood plasma was <0.05 (4 for diclofenac based on previous studies). Ketoprofen only reached up to 0.6M-bM-^@M-0 of the human therapeutic plasma concentration, thus the probability of target-related effects was estimated to be fairly low. Accordingly, a comprehensive analysis of hepatic gene expression revealed no consistent responses. In some contrast, trout exposed to undiluted, treated sewage effluents bioconcentrated ketoprofen and other NSAIDs much more efficiently, according to a meta-analysis of recent studies. Neither of the setups is however an ideal representation of the field situation. If a controlled exposure system with a single chemical in pure water is a reasonable representation of the environment, then the use of ketoprofen is likely to pose a lower risk for wild fish than diclofenac, but if bioconcentration factors from effluent-exposed fish are applied, the risks may be more similar. Ketoprofen (purity M-bM-^IM-%98%) (Sigma-Aldrich, Steinheim, Germany) was dissolved in water (500 mg/L concentrated solution), stirred vigorously and diluted stepwise to obtain stock concentrations (0.5 mg/L, 5 mg/L and 50 mg/L). Juvenile rainbow trout (Oncorhynchus mykiss) of both sexes (age: approximately six months, weight: 39.4M-BM-16.6g) were obtained from VM-CM-$nneM-CM-%ns fiskodling AB, Sweden, and kept in 500 L holding tanks with sand-filtered, aerated fresh water in a flow-through system for one week prior to the experiment. Eight experimental 48 L glass aquaria, semi-covered in black plastic bags to reduce visual stress from the outside (12:12 h light:dark photocycle), were supplied with filtered, aerated freshwater at a flow rate of 0.25 L/min. At the onset of the experiment 80 trout were randomly divided among the aquaria (ten fish in each). Peristaltic pumps (500 M-BM-5l/min) from each stock solution were used to reach target ketoprofen concentrations of 0 (control), 1 (low), 10 (intermediate) and 100 (high) M-BM-5g/L with two aquaria for each concentration. Water samples were taken from each aquarium at day 1, 7 and 14 and stored at -20M-BM-0C until analysis. To minimize biological variation due to variable food intake, the fish were not fed during the exposure period. It should be stressed that trout readily cope with food deprivation for much longer periods at these temperatures [Kullgren et al., 2010]. To ensure a stable water quality, temperature (12.9M-BM-10.1M-BM-0C), pH (7.4M-BM-10.2) and oxygen saturation (88.3%M-BM-11.3) were monitored every third day. The total organic carbon content of the supply water was 2.5 mg/L. At the experiment termination on day 14, the fish were killed by a rapid blow to the head and blood was collected from the caudal vessels by the use of heparinized syringes. The plasma was instantly separated by centrifugation at 10,000 rpm for 2 min and snap frozen in liquid nitrogen. The length and weight of the fish were measured and their sex determined by macroscopical observation of their gonads. Liver samples for gene expression analyses were collected and snap frozen in liquid nitrogen. Both liver and plasma samples were stored at M-bM-^HM-^R70M-BM-0C until analysis. All animal experiments were approved by the local animal committee in Gothenburg (permission number 216-2010). A 15k rainbow trout gene expression microarray was designed for the RT analyzer platform (febit, Heidelberg, Germany) using The Institute for Genomic Research (TIGR) Rainbow Trout Gene Index (RTGI) database version 7.0 (http://compbio.dfci.harvard.edu/tgi/). Details on the probe design strategy, but for eelpout (Zoarces viviparus), and transcript selection strategy are described elsewhere [Cuklev et al., 2011, Kristiansson et al., 2009, Cuklev et al., 2012]. When available, transcripts at GenBank (http://www.ncbi.nlm.nih.gov/nucleotide) were used. Results from similar microarrays, performed by our research group using the same platform, have shown good correlation with qPCR data [Cuklev et al., Kristiansson et al., 2009, Cuklev et al., 2012, Gunnarsson et al., 2009, Lennquist et al., 2011]. The frozen liver tissue was homogenized using Tissuelyser (Qiagen) and total hepatic RNA was isolated using QIAcube and RNeasyM-BM-. Plus Mini Kit (Qiagen). The RNA quantity was determined with spectophotometric measurements (Nanodrop 1000, NanoDrop Technologies) and the quality was assessed by Experion automated electrophoresis using RNA StdSense chip (Bio-Rad, Sundbyberg, Sweden). Biotinylated aRNA was synthesized using MessageAmpM-bM-^DM-" II-Biotin Enhanced Single Round aRNA Amplification Kit (AmbionM-BM-.). The aRNA samples (20 M-BM-5g) were vacuum dried in a vacuum centrifuge, dissolved in 10 M-BM-5l water and fragmented according to the manufacturerM-bM-^@M-^Ys protocol. The following steps described in this subchapter were all performed by febit. Oligonucleotide arrays were synthesized by photo-controlled in situ synthesis using the Geniom One system (febit). Each biochip consists of eight individually accessible micro-channels, each of which is referred to as a microarray in this manuscript. Four individuals (of both sexes) from each control aquarium and each high exposure concentration aquarium were included in the analysis, i.e. eight fish per exposure concentration. In total, two biochips were analyzed corresponding to 16 microarrays. Pre-hybridization and hybridization were performed based on a customized protocol, described in detail by Cuklev et al. [2011]. The samples were randomly allocated on the biochips. Signals were detected using the internal CCD-camera system of the RT analyzer instrument (febit) and quantified using the Geniom Wizard software. Integration times were 266 and 273 ms, determined automatically by the instrument software. All statistical calculations were done in R-2.12.2 (www.r-project.org) [R Development Core Team, 2010]. The quality of pre- and post-normalized arrays was verified with Box and MA plots. The data analysis was performed in the R-package LIMMA [Smyth, 2004]. Data were normalized using the M-bM-^@M-^XquantileM-bM-^@M-^Y method. Moderated t-statistics and adjusted p-values of differential expression were calculated using the empirical Bayes model. As we found no clear differences between the two control aquaria or between the two aquaria with the highest concentration of ketoprofen (100 M-BM-5g/L), all fish from the same treatment were treated as a single group, i.e. we compared one control group to one group of fish exposed to 100 M-BM-5g/L of ketoprofen.

Project description:Background: Diclofenac is a non-steroidal anti-inflammatory drug frequently found in the aquatic environment. Previous studies have reported histological changes in the liver, kidney and gills of fish at concentrations similar to those measured in treated sewage effluents (~1 µg/L). Analyses or predictions of blood plasma levels in fish allow a direct comparison with human therapeutic plasma levels, and may therefore be used to indicate a risk for pharmacological effects in fish. To relate internal exposure to a pharmacological interaction we therefore investigated global hepatic gene expression together with bioconcentration to blood plasma and liver of rainbow trout (Oncorhynchus mykiss) exposed to waterborne diclofenac. Results: At the highest exposure (81.5µg/L) the fish plasma concentration reached ~88% of the human therapeutic levels (Cmax) after two weeks. Using an oligonucleotide microarray followed by quantitative PCR we found extensive effects on hepatic gene expression at this concentration, and some genes were regulated down to the lowest concentration tested (1.6 µg/L) corresponding to ~1.5% of the human Cmax. Functional analysis of differentially expressed genes revealed effects on biological processes such as inflammation and immune response, in agreement with the expected mode of action of diclofenac. In contrast to some previously reported results, the bioconcentration factor was found to be stable (4.02±0.75 for blood plasma and 2.54±0.36 for liver) regardless of the water concentration. Conclusions: Hepatic gene expression is affected in fish at blood levels similar to and below the human Cmax. This adds confidence to the use of fish blood plasma levels to predict risks for pharmacological responses in fish. The microarray results indicate similarities in the mode of action of diclofenac between humans and fish. The identified set of regulated genes may contain useful exposure biomarkers to be used in complex exposure scenarios.

Project description:Effluents from sewage treatment plants contain a mixture of micropollutants with the potential of harming aquatic organisms. Thus, addition of advanced treatment techniques to complement existing conventional methods has been proposed. Some of the advanced techniques could, however, potentially produce additional compounds affecting exposed organisms by unknown modes of action. In the present study the aim was to improve our understanding of how exposure to different sewage effluents affects fish. This was achieved by explorative microarray and quantitative PCR analyses of hepatic gene expression, as well as relative organ sizes of rainbow trout exposed to different sewage effluents (conventionally treated, granular activated carbon, ozonation (5 or 15 mg/L), 5 mg/L ozone plus a moving bed biofilm reactor, or UV-light treatment in combination with hydrogen peroxide). Exposure to the conventionally treated effluent caused a significant increase in liver and heart somatic indexes, an effect removed by all other treatments. Genes connected to xenobiotic metabolism, including cytochrome p450 1A, were differentially expressed in the fish exposed to the conventionally treated effluents, though only effluent treatment with granular activated carbon or ozone at 15 mg/L completely removed this response. The mRNA expression of heat shock protein 70 kDa was induced in all three groups exposed to ozone-treated effluents, suggesting some form of added stress in these fish. The induction of estrogen-responsive genes in the fish exposed to the conventionally treated effluent was effectively reduced by all investigated advanced treatment technologies, although the moving bed biofilm reactor was least efficient. Taken together, granular activated carbon showed the highest potential of reducing responses in fish induced by exposure to sewage effluents. The fish exposure was carried out in March 3-17, 2008 at Henriksdal STP (Stockholm, Sweden) using a large scale pilot plant with parallel treatment lines under ethic permits approved by the local animal committee in Gothenburg (permission no. 36-2007). A thorough description of the different treatment steps (both conventional and advanced) and the experimental setup during the exposure (Fig. 1) has been previously described elsewhere (Samuelsson et al., 2011). In brief, the raw influent underwent conventional activated sludge treatment to generate a reference effluent (Conventional Activated Sludge; CAS). This includes mechanical treatment (removal of large objects using grids with 3 mm column width); primary sedimentation; biological treatment (activated sludge aged 20 days); secondary sedimentation; sand filter. The additional advanced treatment processes were granular activated carbon (GAC), ozonation at 5 mg/L (OZ5) or 15 mg/L (OZ15), ozonation 5 mg/L followed by a moving bed biofilm reactor (MBBR) and irradiation by ultraviolet radiation (0.75 Wh/L) plus hydrogen peroxide (10 mg/L; UH). Two aquaria were used for positive controls. As the tap water available is not as well accepted by the fish, and to better reflect the organic content and overall chemistry of surface water, the control aquaria (TW1 and TW2, together TW) contained carbon-filtered tap water supplemented by 2% of conventionally treated reference effluent. We consider evaluation of implementation and running costs to be outside the scope of this paper. Juvenile rainbow trout (Oncorhynchus mykiss; 113M-BM-123 g; mean bodyweight M-BM-1 standard deviation) of both sexes were obtained from Antens Laxodling AB (AlingsM-CM-%s, Sweden). Rainbow trout is a commonly used species for this type of studies and our lab have ample experience using this species. Furthermore, rainbow trout of sufficient size for metabolomic studies (Samuelsson et al., 2011) can easily be obtained. Finally, the use of trout provides the opportunity for comparisons with other studies (e.g. Gunnarsson et al., 2009 and Albertsson et al., 2010). During an acclimatization period of four days on site, the fish were kept in tanks containing carbon-filtered tap water + 2% reference effluent. On the onset of the experiment, the fish were randomly distributed among eight 100 L aquaria (n=25/aquarium), supplied with treated effluent from the different parallel treatment lines. Aquaria duplicates were used for the controls. During both the acclimatization period as well as the 14 day exposure, the aquaria were aerated and the fish were not fed to reduce variability due to differences in food intake. It should be stressed that trout easily cope with food deprivations for several weeks (Kullgren et al., 2010). Measurements of temperature (13.5M-BM-10.5M-BM-0C), oxygen saturation (97M-BM-13%), pH (7.4M-BM-10.3), conductivity and flow rate were monitored 5 d per week throughout the exposure. In addition, to evaluate differences in removal efficiency between the treatments, total organic carbon (TOC), 7-day biochemical oxygen demand (BOD7), suspended solids (SS), total nitrogen (Tot-N) and total phosphorus (Tot-P) were measured in all effluents (Samuelsson et al., 2011). The sampling took place during two consecutive days (day 14 and 15) due to the large number of fish. For each day, an equal number of fish from each treatment was sampled in random order and killed by a blow to the head. The fish were weighed, their fork length was measured and their sex was determined by macroscopical observation of their gonads. Liver samples were collected for several different studies (hence sample size limitations), frozen on dry ice and stored at -70M-BM-0C until analysis. Homogenization of the frozen liver tissue was carried out using Tissuelyser (Qiagen, Hilden, Germany) and hepatic total RNA was extracted and purified using QIAcube and RNeasyM-BM-. Plus Mini Kit (Qiagen). The RNA quantity and quality were assessed by spectrophotometric measurements with the Nanodrop 1000 (NanoDrop Technologies, Wilmington, DE). To ensure that no degradation had occurred, the isolated RNA was analyzed using Experion automated electrophoresis (Bio-Rad, Hercules, CA). A 15k rainbow trout gene expression microarray was designed for the RT analyzer platform (febit, Heidelberg, Germany) by using The Institute for Genomic Research (TIGR) Rainbow Trout Gene Index (RTGI) database version 7.0 (http://compbio.dfci.harvard.edu/tgi/). Details on the probe design strategy, but for eelpout (Zoarces viviparus), and transcript selection strategy are described elsewhere (Kristiansson et al., 2009; Cuklev et al., 2011). However, in the present study, singletons and non-annotated expressed sequence tags (ESTs) were replaced by newly well-annotated ESTs in rainbow trout. When available, transcripts at GenBank (http://www.ncbi.nlm.nih.gov/nucleotide) were used. In our lab, results from similar microarrays using the same platform have shown good correlation with quantitative polymerase chain reaction (qPCR) data (Gunnarsson et al., 2009b; Kristiansson et al., 2009; Cuklev et al., 2011; Lennquist et al., 2011). To reduce variation in estrogen-responsive genes, only males were used for the subsequent gene expression analyses. Biotinylated antisense RNA (aRNA) was synthesized using MessageAmpM-bM-^DM-" II-Biotin Enhanced Single Round aRNA Amplification Kit (AmbionM-BM-.). The aRNA samples (20 M-BM-5g) were vacuum dried in a vacuum centrifuge, dissolved in 10 M-BM-5l water and fragmented according to the manufacturerM-bM-^@M-^Ys protocol. The following steps described in this subchapter were all performed by febit. Oligonucleotide arrays were synthesized by photo-controlled in situ synthesis using the Geniom One system (febit). Each biochip consists of eight individually accessible micro channels, each of which is referred to as a microarray in this manuscript. Eight individual samples from each aquarium were included in the analysis. In total, 64 microarrays were analyzed. A customized protocol, described in detail elsewhere (Cuklev et al., 2011), was used for prehybridization and hybridization. All samples were randomly placed on the biochips, with one sample from each exposure on each biochip. Signals were detected using the internal CCD-camera system of the RT analyzer instrument (febit) and quantified using the Geniom Wizard software. Integration times were between 156 and 570 ms, determined automatically by the instrument software. All microarray data processing and statistical calculations were performed in R-2.12.2 (www.r-project.org; R Development Core Team, 2010). The quality of pre- and post normalized arrays was verified with box- and MA plots. The data analysis was performed in the R-package LIMMA (Smyth, 2005). Data were normalized using the M-bM-^@M-^XquantileM-bM-^@M-^Y method. Moderated t-statistics and adjusted p-values of differential expression were calculated using the empirical Bayes model.

Project description:Patancheru, near Hyderabad, India, is a major production site for the global bulk drug market. About 90 manufacturers send their wastewater to a common treatment plant in Patancheru. Extraordinary high levels of a wide range of pharmaceuticals have recently been demonstrated in the treated effluent. As little as 0.2% of this effluent can strongly reduce the growth rate of tadpoles, but the underlying mechanisms of toxicity are not known. To begin addressing how the effluent affects aquatic vertebrates, rainbow trout (Oncorhynchus mykiss) were exposed to 0.2% effluent for five days. Several physiological endpoints, together with effects on global hepatic gene expression patterns, were analyzed. The exposed fish showed both an induction of hepatic cytochrome P450 1A (CYP1A) gene expression, as well as EROD activity. Clinical blood chemistry analyses revealed an increase in plasma phosphate levels, which in humans indicates impaired kidney function. Several oxidative stress-related genes were induced in the livers, however, no significant changes in antioxidant enzyme activities or in the hepatic glutathione levels were found. Furthermore, estrogen-regulated genes were slightly up-regulated following exposure, and moderate levels of estriol were detected in the effluent. This study identifies changes in gene expression triggered by exposure to a high dilution of the effluent, supporting the hypothesis that these fish are responding to chemical exposure. The pattern of regulated genes may contribute to the identification of mechanisms of sub-lethal toxicity, as well as illuminate possible causative agents. Two exposure experiments were conducted with control fish and fish exposed to 0.2% effluent. 19 pools of RNA, each containing three fish, were hybrdizied to in situ synthesized custom-made oligonucleotide microarrays. In total, four exposed and four control pools from exposure experiment I and five exposed and six control pools from exposure experiment II were analyzed.

Project description:Atrazine (ATZ) and nonylphenol (NP) are commonly identified contaminants in aquatic habitats; however, relatively few studies have considered the impact of these endocrine disrupters on immune function. This study examined the immunotoxicological effects of ATZ and NP at multiple levels of biological organization. Juvenile rainbow trout (Oncorhynchus mykiss) were exposed to a solvent control (0.00625 % v/v anhydrous ethanol), 59 µg/L ATZ, 555 µg/L ATZ, 2.3 µg/L NP or 18 µg/L NP (measured concentrations) for 4 d. At the end of exposure, fish were assessed for general health indicators (liver somatic index (LSI), spleen somatic index (SSI), hematocrit), biochemical endpoints (plasma cortisol concentrations, lysozyme activity, and protein content), a cellular endpoint (blood leukocyte differential counts), and an integrated immune response at the organism level (host resistance challenge with Listonella anguillarum). Additionally, liver gene expression was assessed using a salmonid microarray (cGRASP, 32K version 1) for fish exposed to the (solvent) control, high ATZ (555 µg/L) and high NP (18 µg/L) treatments. Fish exposed to the high ATZ concentration exhibited elevated plasma cortisol, a decrease in SSI, and decreased lymphocytes and increased monocytes in peripheral blood, with suppression of early immune system processes apparent at the molecular level. In contrast, fish exposed to the high NP concentration showed physiological (e.g. elevated LSI) and gene expression changes (e.g. induction of vitellogenin) consistent with estrogenic effects, as well as decreased lymphocytes in the peripheral blood and more limited alteration in immune system related pathways in the liver transcriptome. Fish exposed to high ATZ or NP concentrations suffered higher mortality than the control group following disease challenge with L. anguillarum. Microarray analysis of the liver transcriptome revealed a total of 211 unique, annotated differentially-regulated genes (DRGs) following high ATZ exposure and 294 DRGs following high NP exposure, with signatures that included alterations to a number of immune-system related processes and pathways. Functional (enrichment) analysis revealed effects on immune system function, metabolism, oxygen homeostasis, cell cycle, DNA damage, and other processes affected by ATZ or NP exposure. Overall this study provides evidence at multiple levels of biological organization that both ATZ and NP are immunotoxic and highlights the potential risk posed by these chemicals to wild fish populations.

Project description:Atrazine (ATZ) and nonylphenol (NP) are commonly identified contaminants in aquatic habitats; however, relatively few studies have considered the impact of these endocrine disrupters on immune function. This study examined the immunotoxicological effects of ATZ and NP at multiple levels of biological organization. Juvenile rainbow trout (Oncorhynchus mykiss) were exposed to a solvent control (0.00625 % v/v anhydrous ethanol), 59 µg/L ATZ, 555 µg/L ATZ, 2.3 µg/L NP or 18 µg/L NP (measured concentrations) for 4 d. At the end of exposure, fish were assessed for general health indicators (liver somatic index (LSI), spleen somatic index (SSI), hematocrit), biochemical endpoints (plasma cortisol concentrations, lysozyme activity, and protein content), a cellular endpoint (blood leukocyte differential counts), and an integrated immune response at the organism level (host resistance challenge with Listonella anguillarum). Additionally, liver gene expression was assessed using a salmonid microarray (cGRASP, 32K version 1) for fish exposed to the (solvent) control, high ATZ (555 µg/L) and high NP (18 µg/L) treatments. Fish exposed to the high ATZ concentration exhibited elevated plasma cortisol, a decrease in SSI, and decreased lymphocytes and increased monocytes in peripheral blood, with suppression of early immune system processes apparent at the molecular level. In contrast, fish exposed to the high NP concentration showed physiological (e.g. elevated LSI) and gene expression changes (e.g. induction of vitellogenin) consistent with estrogenic effects, as well as decreased lymphocytes in the peripheral blood and more limited alteration in immune system related pathways in the liver transcriptome. Fish exposed to high ATZ or NP concentrations suffered higher mortality than the control group following disease challenge with L. anguillarum. Microarray analysis of the liver transcriptome revealed a total of 211 unique, annotated differentially-regulated genes (DRGs) following high ATZ exposure and 294 DRGs following high NP exposure, with signatures that included alterations to a number of immune-system related processes and pathways. Functional (enrichment) analysis revealed effects on immune system function, metabolism, oxygen homeostasis, cell cycle, DNA damage, and other processes affected by ATZ or NP exposure. Overall this study provides evidence at multiple levels of biological organization that both ATZ and NP are immunotoxic and highlights the potential risk posed by these chemicals to wild fish populations. Total of 24 microarrays. 1 experimental sample and 1 pooled reference sample per microarray. Total of 8 individual samples (replicates) per treatment group. Three (3) treatment groups: control, atrazine (ATZ, 555 ug/L) and nonylphenol (NP, 18 ug/L).

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Diclofenac affects global hepatic gene expression in fish at plasma concentrations similar to human therapeutic levels

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