Arch Toxicol (2014) 88:637–646 DOI 10.1007/s00204-013-1186-2
MOLECULAR TOXICOLOGY
Comparison of intake and systemic relative effect potencies of dioxin‑like compounds in female rats after a single oral dose Karin I. van Ede · Patrik L. Andersson · Konrad P. J. Gaisch · Martin van den Berg · Majorie B. M. van Duursen
Received: 18 October 2013 / Accepted: 12 December 2013 / Published online: 21 December 2013 © Springer-Verlag Berlin Heidelberg 2013
Abstract Risk assessment for mixtures of dioxin-like compounds uses the toxic equivalency factor (TEF) approach. Although current WHO-TEFs are mostly based on oral administration, they are commonly used to determine toxicity equivalencies (TEQs) in human blood or tissues. However, the use of “intake” TEFs to calculate systemic TEQs in for example human blood, has never been validated. In this study, intake and systemic relative effect potencies (REPs) for 1,2,3,7,8-pentachlorodibenzo-p-dioxin (PeCDD), 2,3,4,7,8-pentachlorodibenzofuran (4-PeCDF), 3,3′,4,4′,5pentachlorobiphenyl (PCB-126), 2,3′,4,4′,5-pentachlorobiphenyl (PCB-118) and 2,3,3′,4,4′,5-hexachlorobiphenyl (PCB-156) were compared in rats. The effect potencies were calculated based on administered dose and liver, adipose or plasma concentrations in female Sprague–Dawley rats 3 days after a single oral dose, relative to 2,3,7,8-tetrachlorodibenzop-dioxin (TCDD). Hepatic ethoxyresorufin-O-deethylase activity and gene expression of Cyp1a1, 1a2, 1b1 and aryl hydrocarbon receptor repressor in liver and peripheral blood lymphocytes were used as endpoints. Results show that plasma-based systemic REPs were generally within a half log range around the intake REPs for all congeners tested, except for 4-PeCDF. Together with our previously reported
Electronic supplementary material The online version of this article (doi:10.1007/s00204-013-1186-2) contains supplementary material, which is available to authorized users. K. I. van Ede (*) · K. P. J. Gaisch · M. van den Berg · M. B. M. van Duursen Institute for Risk Assessment Sciences, Utrecht University, P.O. Box 80177, 3508 TD Utrecht, The Netherlands e-mail: [email protected] P. L. Andersson Department of Chemistry, Umeå University, Umeå, Sweden
systemic REPs from a mouse study, these data do not warrant the use of systemic REPs as systemic TEFs for human risk assessment. However, further investigation for plasma-based systemic REPs for 4-PeCDF is desirable. Keywords Dibenzofurans · Dioxins · PCBs · Systemic REPs · TEF concept
Introduction Polychlorinated dibenzo-p-dioxins (PCDDs), polychlorinated dibenzofurans (PCDFs) and polychlorinated biphenyls (PCBs) are persistent organic pollutants and commonly occur in the environment and human food chain. Human risk assessment for dioxin-like compounds (DLCs) is challenging because these compounds are present in the environment in complex mixtures. The common approach used by risk assessors is based on the toxic equivalency factor (TEF) concept (Safe 1990, 1994) endorsed by the World Health Organization (WHO) (Van den Berg et al. 1998; Van den Berg et al. 2006). Each congener-specific TEF is an estimate of its relative potency compared to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). In total, 29 PCDDs, PCDFs and dioxin-like PCBs have been assigned with a TEF value. These TEFs are mainly derived from relative effect potencies (REPs) determined in (sub)chronic in vivo studies with the administered dose as exposure metric, resulting in “intake” TEFs (intakeTEFs) (Haws et al. 2006). However, these intakeTEFs are widely used to assess the risk of humans based on concentrations in for example blood. Thereby, it is assumed that an intakeTEF can also be applied for risk assessment when using systemic concentrations in human blood and tissues. However, differences in toxicokinetics may influence the potency of a congener when
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calculated on either administered dose or systemic concentrations (Devito and Birnbaum 1995; Budinsky et al. 2006; DeVito et al. 1997; DeVito et al. 2000). For this reason, the use of intakeTEFs to assess a possible risk based on blood or serum concentrations may potentially lead to a misinterpretation of the actual risk. Currently, there are insufficient data available to either accept or reject the use of intakeTEFs for risk assessment when based on, e.g. blood concentrations (Van den Berg et al. 2006). Previously we described up to one order of magnitude difference between intakeREPs and systemicREPs in C57bl/6 mice for 1,2,3,7,8-pentachlorodibenzo-p-dioxin (PeCDD), 2,3,4,7,8-pentachlorodibenzofuran (4-PeCDF), 3,3′,4,4′,5-pentachlorobiphenyl (PCB-126), 2,3′,4,4′,5-pentachlorobiphenyl (PCB-118) and 2,3,3′,4,4′,5-hexachlorobiphenyl (PCB-156) compared to TCDD, 3 days after a single oral dose (van Ede et al. 2013a). Based on plasma or adipose levels, higher systemicREPs were calculated for PeCDD, 4-PeCDF and PCB-126, and lower systemicREPs for the mono-ortho PCBs 118 and 156 when compared to intakeREPs. In the present study, we describe and compare intakeREPs and systemicREPs for the same congeners in female Sprague–Dawley rats, based on the administered dose or the systemic liver, adipose or plasma concentrations. Similar to our earlier study with mice, intakeREPs and systemic REPs were calculated 3 days after exposure, using sensitive biomarkers for AhR activation, e.g. Cyp1a1, 1a2, 1b1 and aryl hydrocarbon receptor repressor (Ahrr) expression and/or activity in the liver and peripheral blood lymphocytes (PBLs). The results from this study are compared with those from our mouse study and other data from the literature that allow calculations of both intake and systemic REPs.
Materials and methods Chemicals TCDD, PeCDD, 4-PeCDF and PCB-126 were purchased from Wellington Laboratories Inc. (Guelph, Ontario, Canada). After dissolving in corn oil (ACH Food Companies Inc., Oakbrook, IL, USA), concentrations were then checked and confirmed by Wellington Laboratories Inc. PCB-118, PCB-156 and 2,2′,4,4′,5,5′-hexachlorobiphenyl (PCB-153) were purchased from Cerilliant Corp. (Round Rock, TX, USA). These PCBs and corn oil (Sigma-Aldrich, Stockholm, Sweden) were purity checked and, when necessary, purified at the Department of Chemistry, Umeå University, Umeå, Sweden. Final toxicity equivalency (TEQ) contributions of impurities were 6.6 (PCB-118), 36 (PCB156) and 0.41 (PCB-153) ng TEQ/g. These levels were considered to have no influence on the final outcome of
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our results. Further dilutions of the congeners in corn oil (Sigma-Aldrich, Stockholm, Sweden) were prepared at the Institute for Risk Assessment Sciences (IRAS, Utrecht University), the Netherlands. Animals Eight-week-old female Sprague–Dawley rats (Harlan laboratories, Venray, the Netherlands) were randomly assigned to treatment groups (6 animals/group) and allowed to acclimate for 1.5 weeks. The animals were housed in groups in standard cages and conditions (temperature 23 ± 2 °C, 50–60 % relative humidity, 12-h dark and light cycle) with free access to food and water. Rats received a single dose by oral gavage at a dosing volume of 10 ml/kg bw. Depending on the congener used, five different dose levels were administered in the range from 0.5 μg/kg bw (TCDD) up to 500 mg/kg bw (PCB-153), spanning a similar range of administered TEQ doses across congeners based on the 2006 WHO-TEF values. Detailed information regarding the administered dose levels can be found in Supplementary Material; Table S1. Animals were killed 3 days after dosing using CO2/O2. Blood was obtained from the abdominal aorta directly after killing, and liver, thymus, spleen and adipose tissue were removed, weighed (liver and thymus), snap frozen and stored until use at −80 °C. All animal treatments were performed with permission of the Animal Ethical Committee and performed according to Dutch law on Animal Experiments (http://wetten.overheid.nl/BWBR0 003081). Animals were treated humanely and with regard for alleviation of suffering. Compound analysis Analysis of the compounds in blood plasma, adipose and liver tissue samples was performed as described earlier by Van Ede et al. (2013a). In short, adipose and liver tissue samples were cleaned using a combined solid-phase extraction using Na2SO4 and KOH-silica. Blood plasma samples were extracted on an open column using Chem Elut and NaCl. Clean-up was performed using a miniaturized silica column. Samples were spiked after evaporation with 13C-labelled standards. Sample analysis followed the US EPA Method 1613 (http://water.epa.gov/scitech/ methods/cwa/organics/dioxins/index.cfm) using single ion monitoring mode on an Agilent 6809 N (Agilent technologies, Santa Clara, CA, USA) gas chromatograph coupled with a Micromass Ultima Autospec Ultra high resolution mass spectrometer (HRMS, Waters Corp., Milford, MA, USA). To retain unique individual results of each animal, tissue samples (liver, plasma or adipose fat) were not pooled from various animals within the same treatment group but tissues from individual animals that were
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Arch Toxicol (2014) 88:637–646
exposed to different congeners at the same dose level were pooled (TCDD + PeCDD + 4-PeCDF + PCB-126 or PCB-118 + PCB-156 + PCB-153) (See Supplementary Material; Table S1). For example, to determine liver concentrations, a liver from a rat treated with TCDD at the lowest dose was pooled with a liver from another rat treated with PeCDD, one liver from a 4-PeCDF-exposed rat and a liver from a rat dosed with PCB-126 at the lowest dose. This time and cost-effective approach was chosen because a full separation and quantification of individual congeners could be obtained in a single HRGC-HRMS run. The concentrations were calculated on lipid and wet weight basis. Plasma and peripheral blood lymphocyte (PBL) isolation From blood (approximately 7 ml) of each individual rat, plasma and PBLs were isolated using Ficoll-Paque gradient (GE Healthcare Europe, Diegem, Belgium) according to manufacturer’s instructions. Plasma samples were stored directly at −80 °C until compound analysis. Isolated lymphocytes were lysed with RLT buffer (Qiagen, Venlo, the Netherlands) as described in the Qiagen RNAeasy kit protocol and stored until use at −80 °C. EROD activity Hepatic CYP1A1 activity was determined by means of ethoxyresorufin-O-deethylase (EROD) activity in microsomal fractions of liver tissue according to Schulz et al. (2012). RNA isolation and quantitative real‑time polymerase chain reaction (PCR) RNA isolation and quantitative real-time PCR were performed as described earlier by Van Ede et al. (2013a). Primer sequences were as follows: Cyp1a1, forward-5′-ATGTCCA GCTCTCAGATGATAAGGTC-3′ and reverse-5′-ATCCCTG CCAATCACTGTGTCTAAC-3′ (Vondracek et al. 2006), Cyp1a2, forward-5′-GTGAGAACTACAAAGACAACGG TG-3′ and reverse-5′-GTGACTGTTTCAAATCCAGCTC C-3′ (Vondracek et al. 2006), Cyp1b1, forward-5′- CT CATCCTCTTTACCAGATACCCG-3′ and reverse-5′- GA CGTATGGTAAGTTGGGTTGGTC-3′ (Vondracek et al. 2006), Ahrr, forward-5′- CCCCAAGGGGACTTCAGGG GAC-3′ and reverse-5′- TGCTCCAGTCCAGGTGCC TCA-3′ [designed with the Primer designing tool (NCBI)] Arbp, forward-5′- CCTAGAGGGTGTCCGCAATGTG-3′ and reverse-5′- CAGTGGGAAGGTGTAGTCAGTCTC-3′ [designed with the Primer designing tool (NCBI)]. All primers were run through National Center for Biotechnology Information (NCBI) Primer-BLAST database to confirm specificity and validated for optimal annealing temperature (60 °C for all primers) and efficiency. For Cyp1a1,
Cyp1b1, Cyp1a2 and Arbp, the efficiency of the primer pairs was 98–102 % (tested at 60 °C). The Ahrr primer pair efficiency was 120 %. The following programme was used for denaturation and amplification of the cDNA: 3 min at 95 °C, followed by 40 cycles of 15 s at 95 °C and 45 s at 60 °C. Gene expression for each sample was expressed in terms of the threshold cycle (Ct), normalized to the reference gene Arbp (ΔCt). Fold induction was calculated between the treated and vehicle control groups. Data analysis Dose–response curves, effect concentrations and REP calculations for the tested congeners were determined as described previously by Van Ede et al. (2013a). Briefly, dose–response curves were obtained using a sigmoidal dose–response nonlinear regression curve fit with variable slope (GraphPad Prism 6.01, GraphPad Software Inc., San Diego, CA). Next, REPs were calculated using a benchmark response (BMR) approach. To determine REPs, the dose or concentration needed for a congener to reach 20 % of the TCDD response (BMR20TCDD) was calculated. Using the congenerspecific BMR20TCDD concentration, REPs were calculated relatively to TCDD. The selection of the BMR20TCDD concentration instead of effect concentration 50 % (EC50), that generally form the basis of REP determination, was based on several arguments. Many of the obtained dose–response curves in our study did not attain a maximum efficacy or similar Hill slope. Both differences in maximum efficacy and Hill slope could add a significant uncertainty in calculating EC50 values. For this situation, it has been suggested that, e.g. LO(A)ELs or benchmark, dose levels could be used to determine REPs (Van den Berg et al. 2006). In the case of incomplete dose–response curves, also several other studies have suggested the use of other than EC50 values for calculation of REPs (Villeneuve et al. 2000; Toyoshiba et al. 2004; DeVito et al. 2000). The advance of a benchmark approach at the lower part of the dose–response curve is that the lack of agreement in curve shape is less pronounced compared to EC50. Furthermore, in many cases, the BMR20TCDD also present an exposure situation that is more relevant and closer to the actual human exposure. Though the BMR20TCDD value was preferred above a lower BMR value, e.g. BMR10TCDD or BMR05TCDD, as these BMRs would usually fell within the background noise. Thus, REPs were calculated by dividing the concentration of BMR20 of TCDD by the BMR20TCDD concentration of another congener. Statistical analysis Statistically significant differences of the means and variances were determined using analysis of variance (oneway ANOVA) test followed by a Tukey–Kramer multiple
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Arch Toxicol (2014) 88:637–646
comparisons test. Differences were considered statistically significant if P 0.3, a suggested cut-off for liver sequestration (Diliberto et al. 1997), while the mono-ortho PCBs 118, 156, and the nondioxin-like PCB-153 did not show this liver sequestration with liver–adipose concentration ratios of 0.07, 0.13 and 0.06, respectively (Table 1). More details on tissue distribution of these congeners have been described elsewhere by van Ede et al. (2013b). Dose–response curves Dose–response relationships for hepatic EROD activity and gene expression of Cyp1a1, 1b1, 1a2 and Ahrr in liver and PBLs were determined using intake or administered dose levels and liver, adipose tissue or plasma concentrations (See Supplementary Material; Figure S2 and Figure S3). In the liver, all compounds, except the non-dioxin-like PCB-153, significantly induced hepatic EROD activity as well as Cyp1a1, 1b1, 1a2 and Ahrr gene expression. For hepatic EROD activity, TCDD caused already a maximum induction at the lowest dose tested (0.5 μg/kg bw) and it was not possible to define a dose–response curve. Also for PeCDD, 4-PeCDF, PCB-126 and PCB-156, EROD activity was already at 60–75 % of their maximal responses at the lowest doses tested (0.5, 5, 5 and 5,000 μg/kg bw, respectively). A clear distinction in hepatic Cyp1a1, 1b1, 1a2 and Ahrr gene expression was observed between more
13
PCB-126
PCB-118
PCB-156
PCB-153
a
Liver and adipose concentrations (in ng/g tissue) were used to calculate congener-specific ratios. Data represent the mean ± SD of six rats. Statistically significant changes were determined by one-way ANOVA analysis followed by a Tukey’s multiple comparisons test b
Significantly higher from previous concentration (P
MOLECULAR TOXICOLOGY
Comparison of intake and systemic relative effect potencies of dioxin‑like compounds in female rats after a single oral dose Karin I. van Ede · Patrik L. Andersson · Konrad P. J. Gaisch · Martin van den Berg · Majorie B. M. van Duursen
Received: 18 October 2013 / Accepted: 12 December 2013 / Published online: 21 December 2013 © Springer-Verlag Berlin Heidelberg 2013
Abstract Risk assessment for mixtures of dioxin-like compounds uses the toxic equivalency factor (TEF) approach. Although current WHO-TEFs are mostly based on oral administration, they are commonly used to determine toxicity equivalencies (TEQs) in human blood or tissues. However, the use of “intake” TEFs to calculate systemic TEQs in for example human blood, has never been validated. In this study, intake and systemic relative effect potencies (REPs) for 1,2,3,7,8-pentachlorodibenzo-p-dioxin (PeCDD), 2,3,4,7,8-pentachlorodibenzofuran (4-PeCDF), 3,3′,4,4′,5pentachlorobiphenyl (PCB-126), 2,3′,4,4′,5-pentachlorobiphenyl (PCB-118) and 2,3,3′,4,4′,5-hexachlorobiphenyl (PCB-156) were compared in rats. The effect potencies were calculated based on administered dose and liver, adipose or plasma concentrations in female Sprague–Dawley rats 3 days after a single oral dose, relative to 2,3,7,8-tetrachlorodibenzop-dioxin (TCDD). Hepatic ethoxyresorufin-O-deethylase activity and gene expression of Cyp1a1, 1a2, 1b1 and aryl hydrocarbon receptor repressor in liver and peripheral blood lymphocytes were used as endpoints. Results show that plasma-based systemic REPs were generally within a half log range around the intake REPs for all congeners tested, except for 4-PeCDF. Together with our previously reported
Electronic supplementary material The online version of this article (doi:10.1007/s00204-013-1186-2) contains supplementary material, which is available to authorized users. K. I. van Ede (*) · K. P. J. Gaisch · M. van den Berg · M. B. M. van Duursen Institute for Risk Assessment Sciences, Utrecht University, P.O. Box 80177, 3508 TD Utrecht, The Netherlands e-mail: [email protected] P. L. Andersson Department of Chemistry, Umeå University, Umeå, Sweden
systemic REPs from a mouse study, these data do not warrant the use of systemic REPs as systemic TEFs for human risk assessment. However, further investigation for plasma-based systemic REPs for 4-PeCDF is desirable. Keywords Dibenzofurans · Dioxins · PCBs · Systemic REPs · TEF concept
Introduction Polychlorinated dibenzo-p-dioxins (PCDDs), polychlorinated dibenzofurans (PCDFs) and polychlorinated biphenyls (PCBs) are persistent organic pollutants and commonly occur in the environment and human food chain. Human risk assessment for dioxin-like compounds (DLCs) is challenging because these compounds are present in the environment in complex mixtures. The common approach used by risk assessors is based on the toxic equivalency factor (TEF) concept (Safe 1990, 1994) endorsed by the World Health Organization (WHO) (Van den Berg et al. 1998; Van den Berg et al. 2006). Each congener-specific TEF is an estimate of its relative potency compared to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). In total, 29 PCDDs, PCDFs and dioxin-like PCBs have been assigned with a TEF value. These TEFs are mainly derived from relative effect potencies (REPs) determined in (sub)chronic in vivo studies with the administered dose as exposure metric, resulting in “intake” TEFs (intakeTEFs) (Haws et al. 2006). However, these intakeTEFs are widely used to assess the risk of humans based on concentrations in for example blood. Thereby, it is assumed that an intakeTEF can also be applied for risk assessment when using systemic concentrations in human blood and tissues. However, differences in toxicokinetics may influence the potency of a congener when
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calculated on either administered dose or systemic concentrations (Devito and Birnbaum 1995; Budinsky et al. 2006; DeVito et al. 1997; DeVito et al. 2000). For this reason, the use of intakeTEFs to assess a possible risk based on blood or serum concentrations may potentially lead to a misinterpretation of the actual risk. Currently, there are insufficient data available to either accept or reject the use of intakeTEFs for risk assessment when based on, e.g. blood concentrations (Van den Berg et al. 2006). Previously we described up to one order of magnitude difference between intakeREPs and systemicREPs in C57bl/6 mice for 1,2,3,7,8-pentachlorodibenzo-p-dioxin (PeCDD), 2,3,4,7,8-pentachlorodibenzofuran (4-PeCDF), 3,3′,4,4′,5-pentachlorobiphenyl (PCB-126), 2,3′,4,4′,5-pentachlorobiphenyl (PCB-118) and 2,3,3′,4,4′,5-hexachlorobiphenyl (PCB-156) compared to TCDD, 3 days after a single oral dose (van Ede et al. 2013a). Based on plasma or adipose levels, higher systemicREPs were calculated for PeCDD, 4-PeCDF and PCB-126, and lower systemicREPs for the mono-ortho PCBs 118 and 156 when compared to intakeREPs. In the present study, we describe and compare intakeREPs and systemicREPs for the same congeners in female Sprague–Dawley rats, based on the administered dose or the systemic liver, adipose or plasma concentrations. Similar to our earlier study with mice, intakeREPs and systemic REPs were calculated 3 days after exposure, using sensitive biomarkers for AhR activation, e.g. Cyp1a1, 1a2, 1b1 and aryl hydrocarbon receptor repressor (Ahrr) expression and/or activity in the liver and peripheral blood lymphocytes (PBLs). The results from this study are compared with those from our mouse study and other data from the literature that allow calculations of both intake and systemic REPs.
Materials and methods Chemicals TCDD, PeCDD, 4-PeCDF and PCB-126 were purchased from Wellington Laboratories Inc. (Guelph, Ontario, Canada). After dissolving in corn oil (ACH Food Companies Inc., Oakbrook, IL, USA), concentrations were then checked and confirmed by Wellington Laboratories Inc. PCB-118, PCB-156 and 2,2′,4,4′,5,5′-hexachlorobiphenyl (PCB-153) were purchased from Cerilliant Corp. (Round Rock, TX, USA). These PCBs and corn oil (Sigma-Aldrich, Stockholm, Sweden) were purity checked and, when necessary, purified at the Department of Chemistry, Umeå University, Umeå, Sweden. Final toxicity equivalency (TEQ) contributions of impurities were 6.6 (PCB-118), 36 (PCB156) and 0.41 (PCB-153) ng TEQ/g. These levels were considered to have no influence on the final outcome of
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Arch Toxicol (2014) 88:637–646
our results. Further dilutions of the congeners in corn oil (Sigma-Aldrich, Stockholm, Sweden) were prepared at the Institute for Risk Assessment Sciences (IRAS, Utrecht University), the Netherlands. Animals Eight-week-old female Sprague–Dawley rats (Harlan laboratories, Venray, the Netherlands) were randomly assigned to treatment groups (6 animals/group) and allowed to acclimate for 1.5 weeks. The animals were housed in groups in standard cages and conditions (temperature 23 ± 2 °C, 50–60 % relative humidity, 12-h dark and light cycle) with free access to food and water. Rats received a single dose by oral gavage at a dosing volume of 10 ml/kg bw. Depending on the congener used, five different dose levels were administered in the range from 0.5 μg/kg bw (TCDD) up to 500 mg/kg bw (PCB-153), spanning a similar range of administered TEQ doses across congeners based on the 2006 WHO-TEF values. Detailed information regarding the administered dose levels can be found in Supplementary Material; Table S1. Animals were killed 3 days after dosing using CO2/O2. Blood was obtained from the abdominal aorta directly after killing, and liver, thymus, spleen and adipose tissue were removed, weighed (liver and thymus), snap frozen and stored until use at −80 °C. All animal treatments were performed with permission of the Animal Ethical Committee and performed according to Dutch law on Animal Experiments (http://wetten.overheid.nl/BWBR0 003081). Animals were treated humanely and with regard for alleviation of suffering. Compound analysis Analysis of the compounds in blood plasma, adipose and liver tissue samples was performed as described earlier by Van Ede et al. (2013a). In short, adipose and liver tissue samples were cleaned using a combined solid-phase extraction using Na2SO4 and KOH-silica. Blood plasma samples were extracted on an open column using Chem Elut and NaCl. Clean-up was performed using a miniaturized silica column. Samples were spiked after evaporation with 13C-labelled standards. Sample analysis followed the US EPA Method 1613 (http://water.epa.gov/scitech/ methods/cwa/organics/dioxins/index.cfm) using single ion monitoring mode on an Agilent 6809 N (Agilent technologies, Santa Clara, CA, USA) gas chromatograph coupled with a Micromass Ultima Autospec Ultra high resolution mass spectrometer (HRMS, Waters Corp., Milford, MA, USA). To retain unique individual results of each animal, tissue samples (liver, plasma or adipose fat) were not pooled from various animals within the same treatment group but tissues from individual animals that were
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exposed to different congeners at the same dose level were pooled (TCDD + PeCDD + 4-PeCDF + PCB-126 or PCB-118 + PCB-156 + PCB-153) (See Supplementary Material; Table S1). For example, to determine liver concentrations, a liver from a rat treated with TCDD at the lowest dose was pooled with a liver from another rat treated with PeCDD, one liver from a 4-PeCDF-exposed rat and a liver from a rat dosed with PCB-126 at the lowest dose. This time and cost-effective approach was chosen because a full separation and quantification of individual congeners could be obtained in a single HRGC-HRMS run. The concentrations were calculated on lipid and wet weight basis. Plasma and peripheral blood lymphocyte (PBL) isolation From blood (approximately 7 ml) of each individual rat, plasma and PBLs were isolated using Ficoll-Paque gradient (GE Healthcare Europe, Diegem, Belgium) according to manufacturer’s instructions. Plasma samples were stored directly at −80 °C until compound analysis. Isolated lymphocytes were lysed with RLT buffer (Qiagen, Venlo, the Netherlands) as described in the Qiagen RNAeasy kit protocol and stored until use at −80 °C. EROD activity Hepatic CYP1A1 activity was determined by means of ethoxyresorufin-O-deethylase (EROD) activity in microsomal fractions of liver tissue according to Schulz et al. (2012). RNA isolation and quantitative real‑time polymerase chain reaction (PCR) RNA isolation and quantitative real-time PCR were performed as described earlier by Van Ede et al. (2013a). Primer sequences were as follows: Cyp1a1, forward-5′-ATGTCCA GCTCTCAGATGATAAGGTC-3′ and reverse-5′-ATCCCTG CCAATCACTGTGTCTAAC-3′ (Vondracek et al. 2006), Cyp1a2, forward-5′-GTGAGAACTACAAAGACAACGG TG-3′ and reverse-5′-GTGACTGTTTCAAATCCAGCTC C-3′ (Vondracek et al. 2006), Cyp1b1, forward-5′- CT CATCCTCTTTACCAGATACCCG-3′ and reverse-5′- GA CGTATGGTAAGTTGGGTTGGTC-3′ (Vondracek et al. 2006), Ahrr, forward-5′- CCCCAAGGGGACTTCAGGG GAC-3′ and reverse-5′- TGCTCCAGTCCAGGTGCC TCA-3′ [designed with the Primer designing tool (NCBI)] Arbp, forward-5′- CCTAGAGGGTGTCCGCAATGTG-3′ and reverse-5′- CAGTGGGAAGGTGTAGTCAGTCTC-3′ [designed with the Primer designing tool (NCBI)]. All primers were run through National Center for Biotechnology Information (NCBI) Primer-BLAST database to confirm specificity and validated for optimal annealing temperature (60 °C for all primers) and efficiency. For Cyp1a1,
Cyp1b1, Cyp1a2 and Arbp, the efficiency of the primer pairs was 98–102 % (tested at 60 °C). The Ahrr primer pair efficiency was 120 %. The following programme was used for denaturation and amplification of the cDNA: 3 min at 95 °C, followed by 40 cycles of 15 s at 95 °C and 45 s at 60 °C. Gene expression for each sample was expressed in terms of the threshold cycle (Ct), normalized to the reference gene Arbp (ΔCt). Fold induction was calculated between the treated and vehicle control groups. Data analysis Dose–response curves, effect concentrations and REP calculations for the tested congeners were determined as described previously by Van Ede et al. (2013a). Briefly, dose–response curves were obtained using a sigmoidal dose–response nonlinear regression curve fit with variable slope (GraphPad Prism 6.01, GraphPad Software Inc., San Diego, CA). Next, REPs were calculated using a benchmark response (BMR) approach. To determine REPs, the dose or concentration needed for a congener to reach 20 % of the TCDD response (BMR20TCDD) was calculated. Using the congenerspecific BMR20TCDD concentration, REPs were calculated relatively to TCDD. The selection of the BMR20TCDD concentration instead of effect concentration 50 % (EC50), that generally form the basis of REP determination, was based on several arguments. Many of the obtained dose–response curves in our study did not attain a maximum efficacy or similar Hill slope. Both differences in maximum efficacy and Hill slope could add a significant uncertainty in calculating EC50 values. For this situation, it has been suggested that, e.g. LO(A)ELs or benchmark, dose levels could be used to determine REPs (Van den Berg et al. 2006). In the case of incomplete dose–response curves, also several other studies have suggested the use of other than EC50 values for calculation of REPs (Villeneuve et al. 2000; Toyoshiba et al. 2004; DeVito et al. 2000). The advance of a benchmark approach at the lower part of the dose–response curve is that the lack of agreement in curve shape is less pronounced compared to EC50. Furthermore, in many cases, the BMR20TCDD also present an exposure situation that is more relevant and closer to the actual human exposure. Though the BMR20TCDD value was preferred above a lower BMR value, e.g. BMR10TCDD or BMR05TCDD, as these BMRs would usually fell within the background noise. Thus, REPs were calculated by dividing the concentration of BMR20 of TCDD by the BMR20TCDD concentration of another congener. Statistical analysis Statistically significant differences of the means and variances were determined using analysis of variance (oneway ANOVA) test followed by a Tukey–Kramer multiple
13
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Arch Toxicol (2014) 88:637–646
comparisons test. Differences were considered statistically significant if P 0.3, a suggested cut-off for liver sequestration (Diliberto et al. 1997), while the mono-ortho PCBs 118, 156, and the nondioxin-like PCB-153 did not show this liver sequestration with liver–adipose concentration ratios of 0.07, 0.13 and 0.06, respectively (Table 1). More details on tissue distribution of these congeners have been described elsewhere by van Ede et al. (2013b). Dose–response curves Dose–response relationships for hepatic EROD activity and gene expression of Cyp1a1, 1b1, 1a2 and Ahrr in liver and PBLs were determined using intake or administered dose levels and liver, adipose tissue or plasma concentrations (See Supplementary Material; Figure S2 and Figure S3). In the liver, all compounds, except the non-dioxin-like PCB-153, significantly induced hepatic EROD activity as well as Cyp1a1, 1b1, 1a2 and Ahrr gene expression. For hepatic EROD activity, TCDD caused already a maximum induction at the lowest dose tested (0.5 μg/kg bw) and it was not possible to define a dose–response curve. Also for PeCDD, 4-PeCDF, PCB-126 and PCB-156, EROD activity was already at 60–75 % of their maximal responses at the lowest doses tested (0.5, 5, 5 and 5,000 μg/kg bw, respectively). A clear distinction in hepatic Cyp1a1, 1b1, 1a2 and Ahrr gene expression was observed between more
13
PCB-126
PCB-118
PCB-156
PCB-153
a
Liver and adipose concentrations (in ng/g tissue) were used to calculate congener-specific ratios. Data represent the mean ± SD of six rats. Statistically significant changes were determined by one-way ANOVA analysis followed by a Tukey’s multiple comparisons test b
Significantly higher from previous concentration (P
