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Differential detoxifying responses to crude oil water-accommodated fraction in Hyallela curvispina individuals from unpolluted and contaminated sites
Environmental Toxicology and Pharmacology 70 (2019) 103191
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Differential detoxifying responses to crude oil water-accommodated fraction in Hyallela curvispina individuals from unpolluted and contaminated sites
T
J. Del Brioa,b, B.A. Laresa,b, L.B. Parra-Moralesa,b, V.G. Sancheza, C.M. Montagnaa,b, ⁎ A. Venturinoa,c, a
Centro de Investigaciones en Toxicología Ambiental y Agrobiotecnología del Comahue (CITAAC), CONICET, Universidad Nacional del Comahue, Buenos Aires, 1400, Neuquén, Argentina b Facultad de Ciencias del Ambiente y la Salud, Universidad Nacional del Comahue, Buenos Aires, 1400, Neuquén, Argentina c Facultad de Ciencias Agrarias, Universidad Nacional del Comahue, Ruta Nacional 151 12.5 km, Cinco Saltos, Argentina
ARTICLE INFO
ABSTRACT
Keywords: Oil toxicology Amphipods Antioxidant enzymes Cytochrome P450 Biomarkers
Sublethal effects of water-accommodated fraction (WAF) from crude oil of Neuquén basin, Northern PatagoniaArgentina, were examined on both antioxidant and detoxification system of Hyalella curvispina adults collected in Los Barreales (LB) lake and in an oil-polluted stream (DS). The effects of WAF exposure during 6, 24 and 48 h were evaluated in the glutathione content (GSH) and glutathione S-transferase (GST), catalase (CAT) and cytochrome P450 (CYP450) activities. Populations from DS and LB showed not only different basal GSH content and enzyme activities but also different behavior to WAF exposure. LB population exposed to WAF showed a significant increase in GSH content, CAT and CYP450 activities, compared to control group. DS population presented high basal levels in CAT and CYP activity compared with LB population, but their response to WAF exposure was minor. Amphipods from DS, chronically exposed to hydrocarbons, were adapted to their environment.
1. Introduction
of the toxicological research on crude oil contamination has revealed that its toxicity is mainly caused by the water-accommodated fraction (WAF), rather than the dispersed droplets of the oil fraction. Like other xenobiotics, WAF toxic effects on organisms are first shown at the biochemical-molecular level and then transmitted to the following biological levels of organization. In crustaceans, like other organisms, exposure to xenobiotics such as hydrocarbons may result in reactive oxygen species (ROS) production, leading to a condition of oxidative stress (Lavarías et al., 2011; Wu et al., 2011). Oxidative stress is defined as a disturbance in the balance between the production of ROS and antioxidant defenses, in favor of the first ones, which leads to cellular damage at the biomolecule level (Sies, 1993). The cytochrome P450 (CYP450) family are the main enzymes involved in the detoxification of hydrocarbons (Chaty et al., 2004; Rhee et al., 2013). However, ROS can be generated as intermediate metabolites by these oxidative processes. ROS could mediate cellular and extracellular injury via the destruction of membranes, lipids, lipoproteins, or alteration of critical enzyme systems, proteins and ion channels. Organisms have developed an antioxidant defense system in order to minimize oxidative damage to cellular components. This system can be classified into enzymatic and non-enzymatic antioxidants. The first
The development of oil industry and the growing demand for its products have increased the environment vulnerability, especially aquatic ecosystems, to damaging effects of oil pollution (Edema, 2012). Crude oil is a complex mixture of aliphatic and aromatic hydrocarbons along with varying amounts of sulfur, nitrogen, oxygen and trace metals such as nickel, vanadium and chromium (Law, 1993). Water pollution by petroleum hydrocarbons results from accidental damage to pipelines and cisterns, municipal or industrial waste discharges, small daily leakage of navigation and fishing activities, as well as major spills (Pacheco and Santos, 2001). The damage magnitude depends on the site and spilled volume, as well as the permanence time in the affected environment. Generally, aquatic pollution is usually critical in rivers and lakes, because they have relatively low water volumes, unidirectional flows and lower densities than heavier crude oils that lead to sinking phenomena. In addition, oil is retained much longer in lowenergy environments than on wave-swept coasts (Bhattacharyya et al., 2003). Therefore, it is critical to study the impacts of oil pollution on freshwater habitats. Oil contamination may cause serious problems to aquatic life. Most ⁎
Corresponding author at: CITAAC, CONICET, Universidad Nacional del Comahue, Buenos Aires, 1400, Neuquén, Argentina. E-mail address: [email protected] (A. Venturino).
https://doi.org/10.1016/j.etap.2019.04.012 Received 19 December 2018; Received in revised form 26 April 2019; Accepted 30 April 2019 Available online 01 May 2019 1382-6689/ © 2019 Elsevier B.V. All rights reserved.
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group includes oxide-reduction enzymes such as catalase (CAT), superoxide dismutase (SOD), glutathione S-transferases (GST), among others. The second are compounds such as reduced glutathione (GSH), carotenoids, α-tocopherol (vitamin E), ascorbic acid (vitamin C), etc. (Josephy et al., 2005). Among freshwater benthic organisms, amphipods of the species Hyalella curvispina are considered important bioindicators of aquatic ecosystem quality due to their sensitivity to different xenobiotics. In addition, they have been commonly used in toxicological and biochemical studies (Borgmann et al., 2005; Jergentz et al., 2004; Wheelock et al., 2005). Exploitation of energy resources is the main economic activity at Neuquén province in the Northern Patagonia, Argentina. Despite regulations and monitoring programs, hydrocarbons have been detected in water and sediment samples from different sites of Neuquén province (Del Brio et al., 2018; Monza et al., 2013). Previous studies in H. curvispina from both oil-impacted stream and uncontaminated site located in this region, have shown different susceptibility to WAF (Del Brio et al., 2018). The objective of this work was to evaluate the sublethal effects of 6, 24 and 48 h WAF exposure on antioxidant and detoxification system of both H. curvispina populations.
February and March, corresponding to the period of greatest biomass of amphipods. They were kept at a constant temperature of 21 ± 1 °C and a photoperiod regime of 16:8 h L: D with constant aeration. Commercial feed for fish was used ad libitum. Amphipods from LB were kept under laboratory conditions, promoting reproduction. Once a year, the crop was enriched with new individuals, in order to maintain the population. In turn, amphipods from DS were acclimated in the laboratory for two weeks and immediately used in the assays in order to avoid possible modifications in their basal enzymatic activities. 2.4. Amphipod exposures to diluted WAF Freshly prepared WAF was used for amphipods exposure assays. The exposure concentrations of TPHs for each population were chosen to be around their NOEC values, taking in mind that molecular and biochemical detoxifying responses should act below the ecotoxicological levels to be effective (Mardirosian et al., 2017). Amphipods from LB were exposed to a WAF concentration of 0.0014 mg TPH/L, in the range of the NOEC (LC1) previously determined (Del Brio et al., 2018). In turn, as DS amphipods showed no mortality even when exposed to pure WAF, we decided to use an environmental concentration estimated by the geometrical mean between the pure WAF concentration and the maximal TPHs concentration found in DS water (0.35 mg/L) (Del Brio et al., 2018). Control groups were kept in filtered water. The exposure times were 6, 24 and 48 h. Amphipods were held in 1.5 L glass flasks with polypropylene lid to avoid WAF dissipation at constant temperature of 21 ± 1 °C, 8:16 h (L:D) photoperiod and without feeding.
2. Materials and methods 2.1. Chemicals Reduced glutathione (GSH), 7-ethoxycumarine (7-EC), 7-hidroxycoumarine (7-OHC), 1-chloro-2,4-dinitrobenzene (CDNB), 5,5′-dithiobis-2-nitrobenzoic acid (DTNB) and bovine serum albumin were purchased from Sigma Chemical Co., Saint Louis, MO, USA. All the other reagents used were of analytical grade.
2.5. Biochemical assays 2.5.1. Enzyme preparation Groups of ten amphipods were each homogenized in 500 μL of buffer potassium phosphate (143 mM) and EDTA 6.3 mM (pH 7.5) with an electrical homogenizer. The homogenates were centrifuged at 1000 × g for 15 min at 4 °C to eliminate the rests of exoskeleton. From the supernatants, aliquots of 250 and 5 μL were separated for GSH content and protein determination, respectively. The remaining volume of each homogenate was centrifuged at 10,000 x g for 30 min at 4 °C and the supernatants were distributed into aliquots for GST (40 μL), CAT (150 μL) and protein content determinations (40 μL). The GSH content was immediately determined, and the aliquots for GST, CAT and protein content determinations were frozen at −80 °C until biochemical assays were performed. Mean enzyme activity from each group was obtained from 5 independent replicates.
2.2. Preparation of WAF from crude oil The WAF was obtained from crude oil extracted from Chachahuen field, located in Neuquén basin. The fraction was prepared following the methods outlined by Heras et al. (1992), with some modifications. Crude oil was added to filtered freshwater (1:100 v/v) in a 3 L glass flask and homogenized in the dark at room temperature, using a magnetic stirrer at 60 rpm for 24 h. The mixture was then transferred to a separation funnel, where the layers were allowed to separate for 48 h. Finally, the WAF layer was collected and stored at 4 °C. This WAF was mainly composed of C9–C36 n-alkanes, which comprised about 90% of hydrocarbon mass. Other compounds found were single-ring aromatic hydrocarbons (BTEX group: toluene, benzene, xylene, o-xylene and ethylbenzene) and polyaromatic hydrocarbons (PAHs), particularly naphthalene (Del Brio et al., 2018).
2.5.2. GSH content The GSH content was measured by the method of Ellman (1959), with modifications (Venturino et al., 2001). The homogenate was mixed 1:1 v/v with 10% trichloroacetic acid and centrifuged at 10,000 x g for 10 min. The GSH content was immediately measured as acidsoluble thiols in 0.2 mL of supernatant, using 1 mL of 1.5 mM DTNB in 0.25 M potassium phosphate buffer (pH 8.0). The mixture was incubated for 20 min and the absorbance was recorded at 412 nm in an UV/visible spectrophotometer. Acid-soluble thiols were quantified using a standard curve of GSH (1–8 nmol) and expressed as nmol/mg of protein.
2.3. Test species populations Amphipods were collected from two sites, Los Barreales lake (LB) and Durán stream (DS) in Neuquén province. The first one is an artificial lake where petroleum hydrocarbons have not been detected and therefore it was considered as a reference area (S 38° 45′ 34.4″, W 68° 72′ 91.8″). On the other hand, DS originates from an underground interconnection of Limay river ancient course which collect both irrigation and alluvial drainages (S 38° 58′ 20.60″, W 68° 06′ 03.10″). Previous studies have shown that pesticides were below the detection limits. However, contamination with hydrocarbons occurred sometimes into DS by clandestine industrial discharges settled on the margins of the watercourse. Moreover, concentrations of TPHs and PAHs in water and sediment samples from DS has been previously reported (Del Brio et al., 2018; Monza et al., 2013). In addition, DS amphipods were more tolerant to WAF acute and chronic exposure than those from LB (Del Brio et al., 2018). H. curvispina individuals were collected in both sites during
2.5.3. GST activity The GST activity was assayed using CNDB (0.5 mmol/L in acetonitrile) as a substrate (Habig et al., 1974). The reaction mixture in a final volume of 1 mL consisted of 100 mM phosphate buffer (pH 6.5) with 0.5 mM CDNB and 2.5 mM GSH as substrates. The enzymatic reaction was started adding 10 μL of the 10,000 × g supernatant. Absorbance was recorded continuously at 340 nm for 1 min in a UV/visible spectrophotometer at 25 °C (Shimadzu, Kyoto, Japan). Rate measurements were corrected for the non-enzymatic reaction. The activity was 2
Environmental Toxicology and Pharmacology 70 (2019) 103191
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expressed as μmol of CDNB conjugated/min/mg of protein, using a molar extinction coefficient of 9.6 M−1 cm−1.
(p = 0.000004; Fig. 2). The mean GST activity in DS amphipods was 16.81% lower than the one in LB. The exposure to WAF during 6 h did not produce significant differences in GST activity in amphipods from LB compared to control group (p = 0.97). However, there was about 4.10% and 30.52% decreased activity after 24 and 48 h of WAF exposure, respectively (p = 0.037, p = 0.00041). On the other hand, the exposure of DS population to WAF did not significantly affect GST activity at any of the three exposure times (p = 0.62).
2.5.4. CAT activity The CAT activity was determined by recording the decomposition of H2O2 at 240 nm (Beers and Sizer, 1952). The reaction was performed in 1 mL of 50 mM sodium phosphate buffer (pH 7.0) containing 25 mM H2O2 as substrate. The reaction mixture absorbance was controlled to be 1.0 at 240 nm. Forty μL of supernatant were added to initiate the catalyzed reaction, and the absorbance was recorded for 1 min in an UV/visible spectrophotometer at 25 °C (Shimadzu, Kyoto, Japan). Specific activity was expressed as μmol/min/mg of protein using a molar extinction coefficient of 40 M−1 cm−1.
3.3. CAT activity The average of CAT activity in the control group of amphipods from LB (1.18 ± 0.064 μmol/min/mg of protein) was almost 3.5-fold lower than the one from DS (4.11 ± 0.44 μmol/mg of protein) (p = 0.000001; Fig. 3). The CAT activity increased in amphipods from LB exposed to WAF, compared to control group, independently of the exposure times, reaching up 33.74% at 48 h (p = 0.010). In turn, CAT activity significantly decreased at 24 h of exposure to WAF in DS amphipods (p = 0.0021).
2.5.5. CYP450 activity The CYP450 activity was assayed by direct fluorometry using 7-EC as substrate (Bouvier et al., 2002). The assay is based on the O-dealkylation of 7-EC to produce highly fluorescent 7-OHC (ECOD). Amphipods were cut into three fragments and placed altogether in a well containing 50 μL of 50 mM sodium phosphate buffer (pH 7.2). The reaction was initiated by the addition of 50 μL of the developing solution (20 mM 7-EC in 50 mM sodium phosphate buffer, pH 7.2). After 4 h of incubation at 30 °C, the reaction was stopped with 100 μL of a mixture of 0.1 mM glycine buffer pH 10.4/ethanol (1:1). Then, plates were centrifuged for 1 min at 1500 × g to descend the biological tissues. The fluorescence of the samples was measured at 380 nm of excitation and 460 nm of emission in a fluorometer (Hitachi F-7000 Fluorescence Spectrophotometer). A standard curve was measured in every plate with 7-OHC (0.0125–1 nmol) and CYP450 was expressed as pg of 7OHC/adult/min.
3.4. CYP450 activity The average of CYP450 in the control group of amphipods from LB (12.58 ± 0.54 pg 7-OHC/min/adult) was almost 3-fold lower than the one from DS (37.58 ± 1.36 pg 7-OHC/min/adult) (p = 0.000001; Fig. 4). The exposure to WAF significantly increased CYP450 activity with respect to its control in LB population (up to 78%), independently of the exposure time (p = 0.000001). On the other hand, the exposure of DS population to WAF did not significantly affect CYP450 activity at any of the three exposure times (p = 0.16).
2.5.6. Protein determination Protein concentration was determined using bovine serum albumin as the standard curve (5–60 μg) (Lowry et al., 1951).
3.5. Principal Component Analysis (PCA) of the biochemical responses The projection of the individual groups on the component plane showed a sharp separation between LB and DS populations, associated with GSH on one side and with CYP and CAT on the other side of the first component, respectively (Fig. 5A). The first component C1 explained 63.6% of the total variation in the responses. GSH level was negatively correlated with CYP (r=-0.64) and CAT (r=-0.72), while these two enzyme responses were positively correlated between them (r = 0.73). LB groups were further separated by the treatment, with the controls projected close each other on one line over the second component projection R2 axis and the WAF-treated LB individuals spanning into a line to the opposite side of R2. DS treatments showed no apparent grouping pattern respect to the response variables. To further indagate about possible treatment-response variable relationships inside each population, we developed constrained PCAs. For LB population, the four response variables had weight in the two main components explaining in total 98.6% of the variability (Fig. 5B). CYP and GSH were positively correlated (r = 0.99), in the upper C1-C2 positive quadrant. CAT and GST were negatively correlated (r=-095) and in opposite quadrants. Controls in LB were again close each other in a line opposite to the WAF-treated groups that spanned in the sector linked with the responses of CYP, GSH and CAT. For DS population, although CYP and GSH display opposite responses in C1 with a negative correlation of r=0.86, the projection of the different treatments lacks a pattern in the same way as in the global PCA (Fig. 5C).
2.6. Statistical analysis Comparisons within each population and control groups were analyzed by factorial analysis of variance (ANOVA) with Tukey multiple comparison test. When data failed normality (Shapiro-Wilk) or homogeneity of variance (Levene’s test) tests, they were log-transformed (InfoStat, StatSoft Inc. 2007). Principal Components Analysis (PCA) was performed to determine the contribution of the different response variables and the correlations between them, as well as the influence in the separation or aggrupation of the individual factors population (LB or DS), treatment (WAF or controls) and time (6, 24 or 48 h) (NTSYSpc21). 3. Results 3.1. GSH content The average of basal GSH content in the control group of amphipods from LB (5.16 ± 0.44 nmol/mg of protein) was almost 6-fold higher than the one from DS (0.89 ± 0.094 nmol/mg of protein) (p = 0.000001; Fig. 1). WAF exposure significantly increased the GSH content to about 2X with respect to its control in LB population, independently of the exposure time (p = 0.000002). Instead, the GSH content in the DS population increased significantly only after 6 h to WAF exposure (p = 0.00056).
4. Discussion We demonstrate in this work that both populations of H. curvispina present differential detoxifying enzyme patterns in basal conditions. DS amphipods show high levels of CYP450 (ECOD) and CAT activities, compared to LB amphipods. Our PCA (Fig. 5A) supports the association of DS amphipod responses with both enzymes. According with our previously published data (Del Brio et al., 2018), DS amphipods habitat
3.2. GST activity The average GST activity in the control group of amphipods from LB (0.30 ± 0.0066 μmol/min/mg of protein) was significantly different from those collected in DS (0.25 ± 0.0057 μmol/mg of protein) 3
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Fig. 1. GSH contents of H. curvispina from LB and DS exposed to WAF at different times. Each column and bar represent the mean and SE of 5 independent replicates. Asterisks are indicative of significant mean values differences (*** = p < 0.001). Different letters indicate significant differences between basal levels of both populations.
Fig. 2. GST activities of H. curvispina from LB and DS exposed to WAF at different times. Each column and bar represent the mean and SE of 5 independent replicates. Asterisks are indicative of significant mean values differences (* = p < 0.05; *** = p < 0.001). Different letters indicate significant differences between basal levels of both populations.
Fig. 3. CAT activities of H. curvispina from LB and DS exposed to WAF at different times. Each column and bar represent the mean and SE of 5 independent replicates. Asterisks are indicative of significant mean values differences (* = p < 0.05; ** = p < 0.01). Different letters indicate significant differences between basal levels of both populations.
Fig. 4. CYP450 activities of H. curvispina from LB and DS exposed to WAF at different times. Each column and bar represent the mean and SE of 5 independent replicates. Asterisks are indicative of significant mean values differences (*** = p < 0.001). Different letters indicate significant differences between basal levels of both populations.
4
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Fig. 5. Principal Component Analysis (PCA) on biochemical parameter weight in the variability of responses in LB and DS amphipod populations to WAF. A. global analysis including population basal response variability. B. PCA on the variability in LB population. C. PCA on DS population.
presents hydrocarbon contamination from anthropogenic activity in water and sediment. The high basal levels in detoxifying enzymes in DS H. curvispina may be underlying mechanisms of their noticeable resistance to WAF exposure, which is two orders of magnitude or higher compared with LB amphipods inhabiting an uncontaminated site (Del Brio et al., 2018). This differential sensitivity led us initially to study biochemical detoxifying responses at WAF dilutions related to the respective NOEC-population values rather than to a same very low dilution, as we expected that DS population would probably be not responsive at those levels. Nevertheless, once we have determined that some biochemical parameters measured here are potentially good biomarkers, concentration-response experiments for both populations need now to be addressed. The expression of CYP genes can be induced by xenobiotics that bind to the Aryl Hydrocarbon receptor (AhR), which produces an increase in protein synthesis and related enzymatic activity (Honkakoski and Negishi, 2000). Induction of CYP activities has been used as biomarker of exposure to xenobiotics (Martínez-Gómez et al., 2009; Webb, 2011), while overexpression of P450 genes has been attributed to resistance mechanisms (Lindberg et al., 2017). Regarding hydrocarbon contamination, enhanced CYP1A mRNA expression has been demonstrated in liver and gills of rainbow trout exposed to oil spill in a Patagonian stream (Leggieri et al., 2017). In the present study, we demonstrate that CYP450 activity in amphipods from the impacted site is almost 3-fold higher than that of LB. If this effect and/or the lack of further response to WAF exposure are related to AhR-controlled expression, is matter of future studies. We also find that amphipods from the uncontaminated site show an early increasing response of CYP450 activity in the exposure to low WAF concentration. CYP450 activity increases 2-fold at 6 h of exposure to 0.0014 mg TPH/L that is in the NOEC-lethality range for these amphipods. In turn, DS amphipods do not display any induction response in CYP450 activity during the 48 h-exposure to the environmentally relevant TPH concentration of 0.35 mg/L. It is worth mentioning that this concentration is within the NOEC-lethality range of DS amphipods.
The lack of any induction response in CYP450 activity in DS amphipods may be a consequence of the high basal levels impeding further increase. CYP450 induction has been reported in several crustacean species exposed to diverse PAHs such as pyrene and benzo[a]pirene (Oberdörster et al., 2000; Rewitz et al., 2003; Webb, 2011). Similarly, the exposure of crustaceans to WAF induce the expression of genes of the CYP family, which are potential biomarkers of oil contamination (Han et al., 2014). We determined the presence of about 10% of BTEXgroup single aromatic hydrocarbons and naphthalene in the WAF used in H. curvispina exposures (Del Brio et al., 2018). Indeed, the PAHs naphthalene and anthracene were found in DS water and sediments, in the area where DS amphipods were collected (Monza et al., 2013). It is probable that these aromatic hydrocarbons are responsible for the induction responses in CYP450 activity. Our objective is now to determine which CYP genes are specifically being induced, using qPCR technique. GST isoenzymes play a role in activated metabolite conjugation with GSH before excretion, and in protecting tissues from oxidative stress (Kim et al., 2009). Studies on crustaceans exposed to WAF have shown significantly higher GST activities than their control groups (Lavarías et al., 2011; Han et al., 2014). In turn, fish may show decreased GST activities after the exposure to WAF or aromatic hydrocarbon derivatives as a result of oxidative damage (Uguz et al., 2003; Zhang et al., 2004c; Hasselberg et al., 2004; Song et al., 2006). In our study, H. curvispina show little influence of GST activity in the basal differences between LB and DS populations, as remarked by the PCA (Fig. 5A). Only LB amphipods show a relevant decline in GST activity at 48 h of exposure to WAF, but it seems not related to oxidative damage considering that these changes are correlated with CAT activity increase (Fig. 5B). Xenobiotic activation/detoxification through CYP450 is usually associated with oxidative stress from the electron transfer to oxygen atom. The decrease in GSH content because of oxidative stress has been shown in fish exposed to WAF (Zhang et al., 2004a, 2004c; Yin et al., 2007). Slight oxidative stress may induce an increase of GSH synthesis 5
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and detoxifying enzymatic activities, while a severe oxidative stress may cause the oxidation of GSH to GSSG and a decrease in antioxidant enzymes (Elia et al., 2006). The increase in GSH content after hydrocarbons exposure has been reported in aquatic organisms (Hasselberg et al., 2004; Wu et al., 2011), as well as the induction of the antioxidant enzyme CAT (Achuba and Osakwe, 2003; Zhang et al., 2004b; Lavarías et al., 2011; Han et al., 2014; Sandrini-Neto et al., 2016). In the short term, the exposure of LB amphipods to WAF causes an early (6 h) and sustained (up to 48 h) increase in both GSH content and CAT activity. This is in accord with a moderate oxidative stress (Elia et al., 2006). The strong correlation between GSH level and CYP450 activity in LB amphipods (Fig. 5B) suggests that Phase I-detoxification of WAF is causing these moderate levels of oxidative stress. Amphipods from DS, that are chronically exposed to hydrocarbons in their natural habitat, have opposite trends in CYP450 activity and GSH responses, which are negatively correlated (Fig. 5C). The high basal CYP450 activity as an adaptive response and the frequent exposure to hydrocarbons in DS amphipods habitat may be conducing to a substantial oxidative stress situation where GSH is continuously used in ROS breakdown. CAT activity is adaptively high in DS amphipods, contributing to ROS removal. Considering that CAT protein contains a heme group into its active site that is highly sensitive to ROS (Kono and Fridovich, 1982), we may infer that the antioxidant system in DS amphipods is adequately adapted to control oxidative stress. This steadystate of the antioxidant system is reluctant to changes under further exposure of DS amphipods to WAF, complementing the lack of CYP450 response.
American native species Hyalella curvispina. Environ. Toxicol. Pharmacol. 60, 209–215. https://doi.org/10.1016/j.etap.2018.04.022. Edema, N., 2012. Effects of crude oil contaminated water on the environment. In: Manar, E.-S.A.-R. (Ed.), Crude Oil Emulsions Composition Stability and Characterization. InTech, Europe, pp. 169–180. Elia, A.C., Anastasi, V., Dörr, A.J., 2006. Hepatic antioxidant enzymes and total glutathione of Cyprinus carpio exposed to three disinfectants, chlorine dioxide, sodium hypochlorite and peracetic acid, for superficial water potabilization. Chemosphere 64, 1633–1641. https://doi.org/10.1016/j.chemosphere.2006.01.035. Ellman, G.L., 1959. Tissue sulfhydryl groups. Arch. Biochem. Biophys. 82, 70–77. Habig, W.H., Pabst, M.J., Jakoby, W.B., 1974. Glutathione S-transferases. The first enzymatic step in mercapturic acid formation. J. Biol. Chem. 249, 7130–7139. Han, J., Won, E.J., Hwang, D.S., Shin, K.H., Lee, Y.S., Leung, K.M., Lee, S.J., Lee, J.S., 2014. Crude oil exposure results in oxidative stress-mediated dysfunctional development and reproduction in the copepod Tigriopus japonicus and modulates expression of cytochrome P450 (CYP) genes. Aquat. Toxicol. 152, 308–317. https://doi. org/10.1016/j.aquatox.2014.04.027. Hasselberg, L., Meier, S., Svardal, A., 2004. Effects of alkylphenols on redox status in first spawning Atlantic cod (Gadus morhua). Aquat. Toxicol. 69, 95–105. Heras, H., Ackman, R.G., Macpherson, E.J., 1992. Tainting of Atlantic salmon (Salmo salar) by petroleum hydrocarbons during a short term exposure. Mar. Pollut. Bull. 24, 310–315. Honkakoski, P., Negishi, M., 2000. Regulation of cytochrome P450 (CYP) genes by nuclear receptors. Biochem. J. 347, 321–337. Jergentz, S., Pessacq, P., Mugni, H., Bonetto, C., Schulz, R., 2004. Linking in situ bioassays and population dynamics of macroinvertebrates to assess agricultural contamination in streams of the Argentine pampa. Ecotoxicol. Environ. Saf. 59, 133–141. https:// doi.org/10.1016/j.ecoenv.2004.06.007. Josephy, P.D., Guengerich, F.P., Miners, J.O., 2005. Phase I and Phase II" drug metabolism: terminology that we should phase out? Drug Metab. Rev. 37, 575–580. Kim, M., Ahn, I.Y., Cheonand, J., Park, H., 2009. Molecular cloning and thermal stress induced expression of a pi-class glutathione S-transferase (GST) in the Antarctic bivalve Laternula elliptica. Comp. Biochem. Phys. A 152, 207–213. Kono, Y., Fridovich, I., 1982. Superoxide radical inhibits catalase. J. Biol. Chem. 257, 5751–5754. Lavarías, S., Heras, H., Pedrini, N., Tournier, H., Ansaldo, M., 2011. Antioxidant response and oxidative stress levels in Macrobrachium borellii (Crustacea: Palaemonidae) exposed to the water-soluble fraction of petroleum. Comp. Biochem. Physiol. C Toxicol. Pharmacol. 153, 415–421. https://doi.org/10.1016/j.cbpc.2011.02.002. Law, E.A., 1993. Aquatic Pollution: An Introductory Text, second ed. Wiley, New York. Leggieri, L.R., De Anna, J., Cerón, G.A., Arias Darraz, L., Fuentes Monasterio, D., Carcamo, J., Luquet, C., 2017. Cyps molecular biomarkers in rainbow trout Oncorhynchus mykiss to assess oil contamination in a patagonian stream. Rev. Int. Contam. Ambie. 4, 681–690. https://doi.org/10.20937/RICA.2017.33.04.11. Lindberg, C.D., Jayasundara, N., Kozal, J.S., Leuthner, T.C., Di Giulio, R.T., 2017. Resistance to polycyclic aromatic hydrocarbon toxicity and associated bioenergetic consequences in a population of Fundulus heteroclitus. Ecotoxicology 26, 435–448. https://doi.org/10.1007/s10646-017-1775-6. Lowry, O.H., Rosebrough, N.J., Farr, A.L., Randall, R.J., 1951. Protein measurement with the folin phenol reagent. J. Biol. Chem. 193, 265–275. Mardirosian, M.N., Ceschin, D.G., Lascano, C.I., Venturino, A., 2017. Molecular effectors in the chronic exposure to arsenic as early and sensitive biomarkers in developing Rhinella arenarum toads. Aquat. Toxicol. 186, 19–27. https://doi.org/10.1016/j. aquatox.2017.02.019. Martínez-Gómez, C., Fernández, B., Valdés, J., Campillo, J.A., Benedicto, J., Sánchez, F., Vethaak, A.D., 2009. Evaluation of three-year monitoring with biomarkers in fish following the Prestige oil spill (N Spain). Chemosphere 74, 613–620. https://doi.org/ 10.1016/j.chemosphere.2008.10.052. Monza, L.B., Loewy, R.M., Savini, M.C., Pechen de d’Angelo, A.M., 2013. Sources and distribution of aliphatic and polyaromatic hydrocarbons in sediments from the Neuquen River, Argentine Patagonia. J. Environ. Sci. Health Part A Toxic/Hazard. Subst. Environ. Eng. 48, 370–379. https://doi.org/10.1080/10934529.2013.728097. Oberdörster, E., Brouwer, M., Hoexum-Brouwer, T., Manning, S., McLachlan, J.A., 2000. Long-term pyrene exposure of grass shrimp, Palaemonetes pugio, affects molting and reproduction of exposed males and offspring of exposed females. Environ. Health Perspect. 108, 641–646. Pacheco, M., Santos, M.A., 2001. Biotransformation, endocrine, and genetic responses of Anguilla anguilla L. to petroleum distillate products and environmentally contaminated waters. Ecotoxicol. Environ. Saf. 49, 64–75. Rewitz, K., Styrishave, B., Andersen, O., 2003. CYP330A1 and CYP4C39 enzymes in the shore crab Carcinus maenas: sequence and expression regulation by ecdysteroids and xenobiotics. Biochem. Biophys. Res. Commun. 310, 252–260. Rhee, J.S., Kim, B.M., Choi, B.S., Choi, I.Y., Wu, R.S., Nelson, D.R., Lee, J.S., 2013. Whole spectrum of cytochrome P450 genes and molecular responses to water-accommodated fractions exposure in the marine medaka. Environ. Sci. Technol. 47, 4804–4812. https://doi.org/10.1021/es400186r. Sandrini-Neto, L., Pereira, L., Martins, C.C., Silva de Assis, H.C., Camus, L., Lana, P.C., 2016. Antioxidant responses in estuarine invertebrates exposed to repeated oil spills: effects of frequency and dosage in a field manipulative experiment. Aquat. Toxicol. 177, 237–249. https://doi.org/10.1016/j.aquatox.2016.05.028. Sies, H., 1993. Strategies of antioxidant defense. Eur. J. Biochem. 215, 213–219. Song, S.B., Xu, Y., Zhou, B.S., 2006. Effects of hexachlorobenzene on antioxidant status of liver and brain of common carp (Cyprinus carpio). Chemosphere 65, 699–706. https:// doi.org/10.1016/j.chemosphere.2006.01.033. Uguz, C., Iscan, M., Ergüven, A., Isgor, B., Togan, I., 2003. The bioaccumulation of nonyphenol and its adverse effect on the liver of rainbow trout (Onchorynchus
5. Conclusions Amphipods from DS present important differences in the basal levels of antioxidant and detoxification activities compared to LB, probably related to the long-term exposure to hydrocarbons in their habitat. LB amphipods, in turn, show significant responses to short-term WAF exposure in the detoxifying and antioxidant systems. Biochemical parameters, such as CYP450 and CAT activities, could be used as oil pollution biomarkers as a useful tool in environmental monitoring and risk assessment programs. Conflicts of interest The authors have no competing interests to declare. Acknowledgements This work was supported by Grant 04A124 from the Universidad Nacional del Comahue and PICT 2013-1415 from Agencia Nacional de Promoción Científica y Tecnológica. References Achuba, F.I., Osakwe, S.A., 2003. Petroleum-induced free radical toxicity in African catfish (Clarias gariepinus). Fish Physiol. Biochem. 29, 97–103. Beers, R.F., Sizer, I.W., 1952. A spectrophotometric method for measuring the breakdown of hydrogen peroxide by catalase. J. Biol. Chem. 195, 133–140. Bhattacharyya, S., Klerks, P.L., Nyman, J.A., 2003. Toxicity to freshwater organisms from oils and oil spill chemical treatments in laboratory microcosms. Environ. Pollut. 122, 205–215. https://doi.org/10.1016/S0269-7491(02)00294-4. Borgmann, U., Grapentine, L., Norwood, W.P., Bird, G., Dixon, D.G., Lindeman, D., Chapman, P.M., Goulet, R.R., Wang, F., 2005. Sediment toxicity testing with the freshwater amphipod Hyalella azteca: relevance and application. Chemosphere 61, 1740–1745. https://doi.org/10.1016/j.chemosphere.2005.04.011. Bouvier, J.C., Boivin, T., Beslay, D., Sauphanor, B., 2002. Age-dependent response to insecticides and enzymatic variation in susceptible and resistant codling moth larvae. Arch. Insect Biochem. Physiol. 51, 55–66. Chaty, S., Rodius, F., Vasseur, P., 2004. A comparative study of the expression of CYP1A and CYP4 genes in aquatic invertebrate (freshwater mussel, Unio tumidus) and vertebrate (rainbow trout, Oncorhynchus mykiss). Aquat. Toxicol. 69, 81–94. Del Brio, J., Montagna, C.M., Lares, B.A., Parolo, M.E., Venturino, A., 2018. Chemical characterization and toxicity of water-accommodated fraction of oil on the South
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J. Del Brio, et al. mykiss). Environ. Res. 92, 262–270. https://doi.org/10.1016/s0013-9351(03) 00033-1. Venturino, A., Anguiano, O.L., Gauna, L., Cocca, C., Bergoc, R.M., Pechen de D’Angelo, A.M., 2001. Thiols and polyamines in the potentiation of malathion toxicity in larval stages of the toad Bufo arenarum. Comp. Biochem. Physiol. C Toxicol. Pharmacol. 130, 191–198. Webb, D., 2011. Freshwater shrimp (Palaemonetes australis) as a potential bioindicator of crustacean health. Environ. Monit. Assess. 178, 537–544. https://doi.org/10.1007/ s10661-010-1711-1. Wheelock, C.E., Miller, J.L., Miller, M.J., Phillips, B.M., Gee, S.J., Tjeerdema, R.S., Hammock, B.D., 2005. Influence of container adsorption upon observed pyrethroid toxicity to Ceriodaphnia dubia and Hyalella azteca. Aquat. Toxicol. 74, 47–52. https:// doi.org/10.1016/j.aquatox.2005.04.007. Wu, H.Y., Xu, H.X., Hong, Y.B., Zhang, J.F., Wu, J.C., 2011. The use of biomarkers in the antioxidant responses of Daphnia magna to the acute and chronic exposure to no. 20
diesel oil and 2,4-dichlorophenol. Chem. Spec. Bioavailab. 23, 80–87. https://doi. org/10.3184/095422911X13027015402949. Yin, Y., Jia, H., Sun, Y., Yu, H., Wang, X., Wu, J., Xue, Y., 2007. Bioaccumulation and ROS generation in liver of Carassius auratus, exposed to phenanthrene. Comp. Biochem. Physiol. C Toxicol. Pharmacol. 145, 288–293. Zhang, J.F., Shen, H., Wang, X.R., Wu, J.C., Xue, Y.Q., 2004a. Effects of chronic exposure of 2,4-dichlorophenol on the antioxidant system in liver of freshwater goldfish Carassius auratus. Chemosphere 55, 167–174. Zhang, J.F., Sun, Y.Y., Shen, H., Liu, H., Wang, X.R., Wu, J.C., Xue, Y.Q., 2004b. Antioxidant response of Daphnia magna exposed to no. 20 diesel oil. Chem. Spec. Bioavailab. 16, 139–144. Zhang, J.F., Wang, X.R., Guo, H.Y., Wu, J.C., Xue, Y.Q., 2004c. Effects of water-soluble fractions of diesel oil on the antioxidant defenses of the goldfish, Carassius auratus. Ecotoxicol. Environ. Saf. 58, 110–116. https://doi.org/10.1016/j.ecoenv.2003.08. 025.
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Differential detoxifying responses to crude oil water-accommodated fraction in Hyallela curvispina individuals from unpolluted and contaminated sites
T
J. Del Brioa,b, B.A. Laresa,b, L.B. Parra-Moralesa,b, V.G. Sancheza, C.M. Montagnaa,b, ⁎ A. Venturinoa,c, a
Centro de Investigaciones en Toxicología Ambiental y Agrobiotecnología del Comahue (CITAAC), CONICET, Universidad Nacional del Comahue, Buenos Aires, 1400, Neuquén, Argentina b Facultad de Ciencias del Ambiente y la Salud, Universidad Nacional del Comahue, Buenos Aires, 1400, Neuquén, Argentina c Facultad de Ciencias Agrarias, Universidad Nacional del Comahue, Ruta Nacional 151 12.5 km, Cinco Saltos, Argentina
ARTICLE INFO
ABSTRACT
Keywords: Oil toxicology Amphipods Antioxidant enzymes Cytochrome P450 Biomarkers
Sublethal effects of water-accommodated fraction (WAF) from crude oil of Neuquén basin, Northern PatagoniaArgentina, were examined on both antioxidant and detoxification system of Hyalella curvispina adults collected in Los Barreales (LB) lake and in an oil-polluted stream (DS). The effects of WAF exposure during 6, 24 and 48 h were evaluated in the glutathione content (GSH) and glutathione S-transferase (GST), catalase (CAT) and cytochrome P450 (CYP450) activities. Populations from DS and LB showed not only different basal GSH content and enzyme activities but also different behavior to WAF exposure. LB population exposed to WAF showed a significant increase in GSH content, CAT and CYP450 activities, compared to control group. DS population presented high basal levels in CAT and CYP activity compared with LB population, but their response to WAF exposure was minor. Amphipods from DS, chronically exposed to hydrocarbons, were adapted to their environment.
1. Introduction
of the toxicological research on crude oil contamination has revealed that its toxicity is mainly caused by the water-accommodated fraction (WAF), rather than the dispersed droplets of the oil fraction. Like other xenobiotics, WAF toxic effects on organisms are first shown at the biochemical-molecular level and then transmitted to the following biological levels of organization. In crustaceans, like other organisms, exposure to xenobiotics such as hydrocarbons may result in reactive oxygen species (ROS) production, leading to a condition of oxidative stress (Lavarías et al., 2011; Wu et al., 2011). Oxidative stress is defined as a disturbance in the balance between the production of ROS and antioxidant defenses, in favor of the first ones, which leads to cellular damage at the biomolecule level (Sies, 1993). The cytochrome P450 (CYP450) family are the main enzymes involved in the detoxification of hydrocarbons (Chaty et al., 2004; Rhee et al., 2013). However, ROS can be generated as intermediate metabolites by these oxidative processes. ROS could mediate cellular and extracellular injury via the destruction of membranes, lipids, lipoproteins, or alteration of critical enzyme systems, proteins and ion channels. Organisms have developed an antioxidant defense system in order to minimize oxidative damage to cellular components. This system can be classified into enzymatic and non-enzymatic antioxidants. The first
The development of oil industry and the growing demand for its products have increased the environment vulnerability, especially aquatic ecosystems, to damaging effects of oil pollution (Edema, 2012). Crude oil is a complex mixture of aliphatic and aromatic hydrocarbons along with varying amounts of sulfur, nitrogen, oxygen and trace metals such as nickel, vanadium and chromium (Law, 1993). Water pollution by petroleum hydrocarbons results from accidental damage to pipelines and cisterns, municipal or industrial waste discharges, small daily leakage of navigation and fishing activities, as well as major spills (Pacheco and Santos, 2001). The damage magnitude depends on the site and spilled volume, as well as the permanence time in the affected environment. Generally, aquatic pollution is usually critical in rivers and lakes, because they have relatively low water volumes, unidirectional flows and lower densities than heavier crude oils that lead to sinking phenomena. In addition, oil is retained much longer in lowenergy environments than on wave-swept coasts (Bhattacharyya et al., 2003). Therefore, it is critical to study the impacts of oil pollution on freshwater habitats. Oil contamination may cause serious problems to aquatic life. Most ⁎
Corresponding author at: CITAAC, CONICET, Universidad Nacional del Comahue, Buenos Aires, 1400, Neuquén, Argentina. E-mail address: [email protected] (A. Venturino).
https://doi.org/10.1016/j.etap.2019.04.012 Received 19 December 2018; Received in revised form 26 April 2019; Accepted 30 April 2019 Available online 01 May 2019 1382-6689/ © 2019 Elsevier B.V. All rights reserved.
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group includes oxide-reduction enzymes such as catalase (CAT), superoxide dismutase (SOD), glutathione S-transferases (GST), among others. The second are compounds such as reduced glutathione (GSH), carotenoids, α-tocopherol (vitamin E), ascorbic acid (vitamin C), etc. (Josephy et al., 2005). Among freshwater benthic organisms, amphipods of the species Hyalella curvispina are considered important bioindicators of aquatic ecosystem quality due to their sensitivity to different xenobiotics. In addition, they have been commonly used in toxicological and biochemical studies (Borgmann et al., 2005; Jergentz et al., 2004; Wheelock et al., 2005). Exploitation of energy resources is the main economic activity at Neuquén province in the Northern Patagonia, Argentina. Despite regulations and monitoring programs, hydrocarbons have been detected in water and sediment samples from different sites of Neuquén province (Del Brio et al., 2018; Monza et al., 2013). Previous studies in H. curvispina from both oil-impacted stream and uncontaminated site located in this region, have shown different susceptibility to WAF (Del Brio et al., 2018). The objective of this work was to evaluate the sublethal effects of 6, 24 and 48 h WAF exposure on antioxidant and detoxification system of both H. curvispina populations.
February and March, corresponding to the period of greatest biomass of amphipods. They were kept at a constant temperature of 21 ± 1 °C and a photoperiod regime of 16:8 h L: D with constant aeration. Commercial feed for fish was used ad libitum. Amphipods from LB were kept under laboratory conditions, promoting reproduction. Once a year, the crop was enriched with new individuals, in order to maintain the population. In turn, amphipods from DS were acclimated in the laboratory for two weeks and immediately used in the assays in order to avoid possible modifications in their basal enzymatic activities. 2.4. Amphipod exposures to diluted WAF Freshly prepared WAF was used for amphipods exposure assays. The exposure concentrations of TPHs for each population were chosen to be around their NOEC values, taking in mind that molecular and biochemical detoxifying responses should act below the ecotoxicological levels to be effective (Mardirosian et al., 2017). Amphipods from LB were exposed to a WAF concentration of 0.0014 mg TPH/L, in the range of the NOEC (LC1) previously determined (Del Brio et al., 2018). In turn, as DS amphipods showed no mortality even when exposed to pure WAF, we decided to use an environmental concentration estimated by the geometrical mean between the pure WAF concentration and the maximal TPHs concentration found in DS water (0.35 mg/L) (Del Brio et al., 2018). Control groups were kept in filtered water. The exposure times were 6, 24 and 48 h. Amphipods were held in 1.5 L glass flasks with polypropylene lid to avoid WAF dissipation at constant temperature of 21 ± 1 °C, 8:16 h (L:D) photoperiod and without feeding.
2. Materials and methods 2.1. Chemicals Reduced glutathione (GSH), 7-ethoxycumarine (7-EC), 7-hidroxycoumarine (7-OHC), 1-chloro-2,4-dinitrobenzene (CDNB), 5,5′-dithiobis-2-nitrobenzoic acid (DTNB) and bovine serum albumin were purchased from Sigma Chemical Co., Saint Louis, MO, USA. All the other reagents used were of analytical grade.
2.5. Biochemical assays 2.5.1. Enzyme preparation Groups of ten amphipods were each homogenized in 500 μL of buffer potassium phosphate (143 mM) and EDTA 6.3 mM (pH 7.5) with an electrical homogenizer. The homogenates were centrifuged at 1000 × g for 15 min at 4 °C to eliminate the rests of exoskeleton. From the supernatants, aliquots of 250 and 5 μL were separated for GSH content and protein determination, respectively. The remaining volume of each homogenate was centrifuged at 10,000 x g for 30 min at 4 °C and the supernatants were distributed into aliquots for GST (40 μL), CAT (150 μL) and protein content determinations (40 μL). The GSH content was immediately determined, and the aliquots for GST, CAT and protein content determinations were frozen at −80 °C until biochemical assays were performed. Mean enzyme activity from each group was obtained from 5 independent replicates.
2.2. Preparation of WAF from crude oil The WAF was obtained from crude oil extracted from Chachahuen field, located in Neuquén basin. The fraction was prepared following the methods outlined by Heras et al. (1992), with some modifications. Crude oil was added to filtered freshwater (1:100 v/v) in a 3 L glass flask and homogenized in the dark at room temperature, using a magnetic stirrer at 60 rpm for 24 h. The mixture was then transferred to a separation funnel, where the layers were allowed to separate for 48 h. Finally, the WAF layer was collected and stored at 4 °C. This WAF was mainly composed of C9–C36 n-alkanes, which comprised about 90% of hydrocarbon mass. Other compounds found were single-ring aromatic hydrocarbons (BTEX group: toluene, benzene, xylene, o-xylene and ethylbenzene) and polyaromatic hydrocarbons (PAHs), particularly naphthalene (Del Brio et al., 2018).
2.5.2. GSH content The GSH content was measured by the method of Ellman (1959), with modifications (Venturino et al., 2001). The homogenate was mixed 1:1 v/v with 10% trichloroacetic acid and centrifuged at 10,000 x g for 10 min. The GSH content was immediately measured as acidsoluble thiols in 0.2 mL of supernatant, using 1 mL of 1.5 mM DTNB in 0.25 M potassium phosphate buffer (pH 8.0). The mixture was incubated for 20 min and the absorbance was recorded at 412 nm in an UV/visible spectrophotometer. Acid-soluble thiols were quantified using a standard curve of GSH (1–8 nmol) and expressed as nmol/mg of protein.
2.3. Test species populations Amphipods were collected from two sites, Los Barreales lake (LB) and Durán stream (DS) in Neuquén province. The first one is an artificial lake where petroleum hydrocarbons have not been detected and therefore it was considered as a reference area (S 38° 45′ 34.4″, W 68° 72′ 91.8″). On the other hand, DS originates from an underground interconnection of Limay river ancient course which collect both irrigation and alluvial drainages (S 38° 58′ 20.60″, W 68° 06′ 03.10″). Previous studies have shown that pesticides were below the detection limits. However, contamination with hydrocarbons occurred sometimes into DS by clandestine industrial discharges settled on the margins of the watercourse. Moreover, concentrations of TPHs and PAHs in water and sediment samples from DS has been previously reported (Del Brio et al., 2018; Monza et al., 2013). In addition, DS amphipods were more tolerant to WAF acute and chronic exposure than those from LB (Del Brio et al., 2018). H. curvispina individuals were collected in both sites during
2.5.3. GST activity The GST activity was assayed using CNDB (0.5 mmol/L in acetonitrile) as a substrate (Habig et al., 1974). The reaction mixture in a final volume of 1 mL consisted of 100 mM phosphate buffer (pH 6.5) with 0.5 mM CDNB and 2.5 mM GSH as substrates. The enzymatic reaction was started adding 10 μL of the 10,000 × g supernatant. Absorbance was recorded continuously at 340 nm for 1 min in a UV/visible spectrophotometer at 25 °C (Shimadzu, Kyoto, Japan). Rate measurements were corrected for the non-enzymatic reaction. The activity was 2
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expressed as μmol of CDNB conjugated/min/mg of protein, using a molar extinction coefficient of 9.6 M−1 cm−1.
(p = 0.000004; Fig. 2). The mean GST activity in DS amphipods was 16.81% lower than the one in LB. The exposure to WAF during 6 h did not produce significant differences in GST activity in amphipods from LB compared to control group (p = 0.97). However, there was about 4.10% and 30.52% decreased activity after 24 and 48 h of WAF exposure, respectively (p = 0.037, p = 0.00041). On the other hand, the exposure of DS population to WAF did not significantly affect GST activity at any of the three exposure times (p = 0.62).
2.5.4. CAT activity The CAT activity was determined by recording the decomposition of H2O2 at 240 nm (Beers and Sizer, 1952). The reaction was performed in 1 mL of 50 mM sodium phosphate buffer (pH 7.0) containing 25 mM H2O2 as substrate. The reaction mixture absorbance was controlled to be 1.0 at 240 nm. Forty μL of supernatant were added to initiate the catalyzed reaction, and the absorbance was recorded for 1 min in an UV/visible spectrophotometer at 25 °C (Shimadzu, Kyoto, Japan). Specific activity was expressed as μmol/min/mg of protein using a molar extinction coefficient of 40 M−1 cm−1.
3.3. CAT activity The average of CAT activity in the control group of amphipods from LB (1.18 ± 0.064 μmol/min/mg of protein) was almost 3.5-fold lower than the one from DS (4.11 ± 0.44 μmol/mg of protein) (p = 0.000001; Fig. 3). The CAT activity increased in amphipods from LB exposed to WAF, compared to control group, independently of the exposure times, reaching up 33.74% at 48 h (p = 0.010). In turn, CAT activity significantly decreased at 24 h of exposure to WAF in DS amphipods (p = 0.0021).
2.5.5. CYP450 activity The CYP450 activity was assayed by direct fluorometry using 7-EC as substrate (Bouvier et al., 2002). The assay is based on the O-dealkylation of 7-EC to produce highly fluorescent 7-OHC (ECOD). Amphipods were cut into three fragments and placed altogether in a well containing 50 μL of 50 mM sodium phosphate buffer (pH 7.2). The reaction was initiated by the addition of 50 μL of the developing solution (20 mM 7-EC in 50 mM sodium phosphate buffer, pH 7.2). After 4 h of incubation at 30 °C, the reaction was stopped with 100 μL of a mixture of 0.1 mM glycine buffer pH 10.4/ethanol (1:1). Then, plates were centrifuged for 1 min at 1500 × g to descend the biological tissues. The fluorescence of the samples was measured at 380 nm of excitation and 460 nm of emission in a fluorometer (Hitachi F-7000 Fluorescence Spectrophotometer). A standard curve was measured in every plate with 7-OHC (0.0125–1 nmol) and CYP450 was expressed as pg of 7OHC/adult/min.
3.4. CYP450 activity The average of CYP450 in the control group of amphipods from LB (12.58 ± 0.54 pg 7-OHC/min/adult) was almost 3-fold lower than the one from DS (37.58 ± 1.36 pg 7-OHC/min/adult) (p = 0.000001; Fig. 4). The exposure to WAF significantly increased CYP450 activity with respect to its control in LB population (up to 78%), independently of the exposure time (p = 0.000001). On the other hand, the exposure of DS population to WAF did not significantly affect CYP450 activity at any of the three exposure times (p = 0.16).
2.5.6. Protein determination Protein concentration was determined using bovine serum albumin as the standard curve (5–60 μg) (Lowry et al., 1951).
3.5. Principal Component Analysis (PCA) of the biochemical responses The projection of the individual groups on the component plane showed a sharp separation between LB and DS populations, associated with GSH on one side and with CYP and CAT on the other side of the first component, respectively (Fig. 5A). The first component C1 explained 63.6% of the total variation in the responses. GSH level was negatively correlated with CYP (r=-0.64) and CAT (r=-0.72), while these two enzyme responses were positively correlated between them (r = 0.73). LB groups were further separated by the treatment, with the controls projected close each other on one line over the second component projection R2 axis and the WAF-treated LB individuals spanning into a line to the opposite side of R2. DS treatments showed no apparent grouping pattern respect to the response variables. To further indagate about possible treatment-response variable relationships inside each population, we developed constrained PCAs. For LB population, the four response variables had weight in the two main components explaining in total 98.6% of the variability (Fig. 5B). CYP and GSH were positively correlated (r = 0.99), in the upper C1-C2 positive quadrant. CAT and GST were negatively correlated (r=-095) and in opposite quadrants. Controls in LB were again close each other in a line opposite to the WAF-treated groups that spanned in the sector linked with the responses of CYP, GSH and CAT. For DS population, although CYP and GSH display opposite responses in C1 with a negative correlation of r=0.86, the projection of the different treatments lacks a pattern in the same way as in the global PCA (Fig. 5C).
2.6. Statistical analysis Comparisons within each population and control groups were analyzed by factorial analysis of variance (ANOVA) with Tukey multiple comparison test. When data failed normality (Shapiro-Wilk) or homogeneity of variance (Levene’s test) tests, they were log-transformed (InfoStat, StatSoft Inc. 2007). Principal Components Analysis (PCA) was performed to determine the contribution of the different response variables and the correlations between them, as well as the influence in the separation or aggrupation of the individual factors population (LB or DS), treatment (WAF or controls) and time (6, 24 or 48 h) (NTSYSpc21). 3. Results 3.1. GSH content The average of basal GSH content in the control group of amphipods from LB (5.16 ± 0.44 nmol/mg of protein) was almost 6-fold higher than the one from DS (0.89 ± 0.094 nmol/mg of protein) (p = 0.000001; Fig. 1). WAF exposure significantly increased the GSH content to about 2X with respect to its control in LB population, independently of the exposure time (p = 0.000002). Instead, the GSH content in the DS population increased significantly only after 6 h to WAF exposure (p = 0.00056).
4. Discussion We demonstrate in this work that both populations of H. curvispina present differential detoxifying enzyme patterns in basal conditions. DS amphipods show high levels of CYP450 (ECOD) and CAT activities, compared to LB amphipods. Our PCA (Fig. 5A) supports the association of DS amphipod responses with both enzymes. According with our previously published data (Del Brio et al., 2018), DS amphipods habitat
3.2. GST activity The average GST activity in the control group of amphipods from LB (0.30 ± 0.0066 μmol/min/mg of protein) was significantly different from those collected in DS (0.25 ± 0.0057 μmol/mg of protein) 3
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Fig. 1. GSH contents of H. curvispina from LB and DS exposed to WAF at different times. Each column and bar represent the mean and SE of 5 independent replicates. Asterisks are indicative of significant mean values differences (*** = p < 0.001). Different letters indicate significant differences between basal levels of both populations.
Fig. 2. GST activities of H. curvispina from LB and DS exposed to WAF at different times. Each column and bar represent the mean and SE of 5 independent replicates. Asterisks are indicative of significant mean values differences (* = p < 0.05; *** = p < 0.001). Different letters indicate significant differences between basal levels of both populations.
Fig. 3. CAT activities of H. curvispina from LB and DS exposed to WAF at different times. Each column and bar represent the mean and SE of 5 independent replicates. Asterisks are indicative of significant mean values differences (* = p < 0.05; ** = p < 0.01). Different letters indicate significant differences between basal levels of both populations.
Fig. 4. CYP450 activities of H. curvispina from LB and DS exposed to WAF at different times. Each column and bar represent the mean and SE of 5 independent replicates. Asterisks are indicative of significant mean values differences (*** = p < 0.001). Different letters indicate significant differences between basal levels of both populations.
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Fig. 5. Principal Component Analysis (PCA) on biochemical parameter weight in the variability of responses in LB and DS amphipod populations to WAF. A. global analysis including population basal response variability. B. PCA on the variability in LB population. C. PCA on DS population.
presents hydrocarbon contamination from anthropogenic activity in water and sediment. The high basal levels in detoxifying enzymes in DS H. curvispina may be underlying mechanisms of their noticeable resistance to WAF exposure, which is two orders of magnitude or higher compared with LB amphipods inhabiting an uncontaminated site (Del Brio et al., 2018). This differential sensitivity led us initially to study biochemical detoxifying responses at WAF dilutions related to the respective NOEC-population values rather than to a same very low dilution, as we expected that DS population would probably be not responsive at those levels. Nevertheless, once we have determined that some biochemical parameters measured here are potentially good biomarkers, concentration-response experiments for both populations need now to be addressed. The expression of CYP genes can be induced by xenobiotics that bind to the Aryl Hydrocarbon receptor (AhR), which produces an increase in protein synthesis and related enzymatic activity (Honkakoski and Negishi, 2000). Induction of CYP activities has been used as biomarker of exposure to xenobiotics (Martínez-Gómez et al., 2009; Webb, 2011), while overexpression of P450 genes has been attributed to resistance mechanisms (Lindberg et al., 2017). Regarding hydrocarbon contamination, enhanced CYP1A mRNA expression has been demonstrated in liver and gills of rainbow trout exposed to oil spill in a Patagonian stream (Leggieri et al., 2017). In the present study, we demonstrate that CYP450 activity in amphipods from the impacted site is almost 3-fold higher than that of LB. If this effect and/or the lack of further response to WAF exposure are related to AhR-controlled expression, is matter of future studies. We also find that amphipods from the uncontaminated site show an early increasing response of CYP450 activity in the exposure to low WAF concentration. CYP450 activity increases 2-fold at 6 h of exposure to 0.0014 mg TPH/L that is in the NOEC-lethality range for these amphipods. In turn, DS amphipods do not display any induction response in CYP450 activity during the 48 h-exposure to the environmentally relevant TPH concentration of 0.35 mg/L. It is worth mentioning that this concentration is within the NOEC-lethality range of DS amphipods.
The lack of any induction response in CYP450 activity in DS amphipods may be a consequence of the high basal levels impeding further increase. CYP450 induction has been reported in several crustacean species exposed to diverse PAHs such as pyrene and benzo[a]pirene (Oberdörster et al., 2000; Rewitz et al., 2003; Webb, 2011). Similarly, the exposure of crustaceans to WAF induce the expression of genes of the CYP family, which are potential biomarkers of oil contamination (Han et al., 2014). We determined the presence of about 10% of BTEXgroup single aromatic hydrocarbons and naphthalene in the WAF used in H. curvispina exposures (Del Brio et al., 2018). Indeed, the PAHs naphthalene and anthracene were found in DS water and sediments, in the area where DS amphipods were collected (Monza et al., 2013). It is probable that these aromatic hydrocarbons are responsible for the induction responses in CYP450 activity. Our objective is now to determine which CYP genes are specifically being induced, using qPCR technique. GST isoenzymes play a role in activated metabolite conjugation with GSH before excretion, and in protecting tissues from oxidative stress (Kim et al., 2009). Studies on crustaceans exposed to WAF have shown significantly higher GST activities than their control groups (Lavarías et al., 2011; Han et al., 2014). In turn, fish may show decreased GST activities after the exposure to WAF or aromatic hydrocarbon derivatives as a result of oxidative damage (Uguz et al., 2003; Zhang et al., 2004c; Hasselberg et al., 2004; Song et al., 2006). In our study, H. curvispina show little influence of GST activity in the basal differences between LB and DS populations, as remarked by the PCA (Fig. 5A). Only LB amphipods show a relevant decline in GST activity at 48 h of exposure to WAF, but it seems not related to oxidative damage considering that these changes are correlated with CAT activity increase (Fig. 5B). Xenobiotic activation/detoxification through CYP450 is usually associated with oxidative stress from the electron transfer to oxygen atom. The decrease in GSH content because of oxidative stress has been shown in fish exposed to WAF (Zhang et al., 2004a, 2004c; Yin et al., 2007). Slight oxidative stress may induce an increase of GSH synthesis 5
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and detoxifying enzymatic activities, while a severe oxidative stress may cause the oxidation of GSH to GSSG and a decrease in antioxidant enzymes (Elia et al., 2006). The increase in GSH content after hydrocarbons exposure has been reported in aquatic organisms (Hasselberg et al., 2004; Wu et al., 2011), as well as the induction of the antioxidant enzyme CAT (Achuba and Osakwe, 2003; Zhang et al., 2004b; Lavarías et al., 2011; Han et al., 2014; Sandrini-Neto et al., 2016). In the short term, the exposure of LB amphipods to WAF causes an early (6 h) and sustained (up to 48 h) increase in both GSH content and CAT activity. This is in accord with a moderate oxidative stress (Elia et al., 2006). The strong correlation between GSH level and CYP450 activity in LB amphipods (Fig. 5B) suggests that Phase I-detoxification of WAF is causing these moderate levels of oxidative stress. Amphipods from DS, that are chronically exposed to hydrocarbons in their natural habitat, have opposite trends in CYP450 activity and GSH responses, which are negatively correlated (Fig. 5C). The high basal CYP450 activity as an adaptive response and the frequent exposure to hydrocarbons in DS amphipods habitat may be conducing to a substantial oxidative stress situation where GSH is continuously used in ROS breakdown. CAT activity is adaptively high in DS amphipods, contributing to ROS removal. Considering that CAT protein contains a heme group into its active site that is highly sensitive to ROS (Kono and Fridovich, 1982), we may infer that the antioxidant system in DS amphipods is adequately adapted to control oxidative stress. This steadystate of the antioxidant system is reluctant to changes under further exposure of DS amphipods to WAF, complementing the lack of CYP450 response.
American native species Hyalella curvispina. Environ. Toxicol. Pharmacol. 60, 209–215. https://doi.org/10.1016/j.etap.2018.04.022. Edema, N., 2012. Effects of crude oil contaminated water on the environment. In: Manar, E.-S.A.-R. (Ed.), Crude Oil Emulsions Composition Stability and Characterization. InTech, Europe, pp. 169–180. Elia, A.C., Anastasi, V., Dörr, A.J., 2006. Hepatic antioxidant enzymes and total glutathione of Cyprinus carpio exposed to three disinfectants, chlorine dioxide, sodium hypochlorite and peracetic acid, for superficial water potabilization. Chemosphere 64, 1633–1641. https://doi.org/10.1016/j.chemosphere.2006.01.035. Ellman, G.L., 1959. Tissue sulfhydryl groups. Arch. Biochem. Biophys. 82, 70–77. Habig, W.H., Pabst, M.J., Jakoby, W.B., 1974. Glutathione S-transferases. The first enzymatic step in mercapturic acid formation. J. Biol. Chem. 249, 7130–7139. Han, J., Won, E.J., Hwang, D.S., Shin, K.H., Lee, Y.S., Leung, K.M., Lee, S.J., Lee, J.S., 2014. Crude oil exposure results in oxidative stress-mediated dysfunctional development and reproduction in the copepod Tigriopus japonicus and modulates expression of cytochrome P450 (CYP) genes. Aquat. Toxicol. 152, 308–317. https://doi. org/10.1016/j.aquatox.2014.04.027. Hasselberg, L., Meier, S., Svardal, A., 2004. Effects of alkylphenols on redox status in first spawning Atlantic cod (Gadus morhua). Aquat. Toxicol. 69, 95–105. Heras, H., Ackman, R.G., Macpherson, E.J., 1992. Tainting of Atlantic salmon (Salmo salar) by petroleum hydrocarbons during a short term exposure. Mar. Pollut. Bull. 24, 310–315. Honkakoski, P., Negishi, M., 2000. Regulation of cytochrome P450 (CYP) genes by nuclear receptors. Biochem. J. 347, 321–337. Jergentz, S., Pessacq, P., Mugni, H., Bonetto, C., Schulz, R., 2004. Linking in situ bioassays and population dynamics of macroinvertebrates to assess agricultural contamination in streams of the Argentine pampa. Ecotoxicol. Environ. Saf. 59, 133–141. https:// doi.org/10.1016/j.ecoenv.2004.06.007. Josephy, P.D., Guengerich, F.P., Miners, J.O., 2005. Phase I and Phase II" drug metabolism: terminology that we should phase out? Drug Metab. Rev. 37, 575–580. Kim, M., Ahn, I.Y., Cheonand, J., Park, H., 2009. Molecular cloning and thermal stress induced expression of a pi-class glutathione S-transferase (GST) in the Antarctic bivalve Laternula elliptica. Comp. Biochem. Phys. A 152, 207–213. Kono, Y., Fridovich, I., 1982. Superoxide radical inhibits catalase. J. Biol. Chem. 257, 5751–5754. Lavarías, S., Heras, H., Pedrini, N., Tournier, H., Ansaldo, M., 2011. Antioxidant response and oxidative stress levels in Macrobrachium borellii (Crustacea: Palaemonidae) exposed to the water-soluble fraction of petroleum. Comp. Biochem. Physiol. C Toxicol. Pharmacol. 153, 415–421. https://doi.org/10.1016/j.cbpc.2011.02.002. Law, E.A., 1993. Aquatic Pollution: An Introductory Text, second ed. Wiley, New York. Leggieri, L.R., De Anna, J., Cerón, G.A., Arias Darraz, L., Fuentes Monasterio, D., Carcamo, J., Luquet, C., 2017. Cyps molecular biomarkers in rainbow trout Oncorhynchus mykiss to assess oil contamination in a patagonian stream. Rev. Int. Contam. Ambie. 4, 681–690. https://doi.org/10.20937/RICA.2017.33.04.11. Lindberg, C.D., Jayasundara, N., Kozal, J.S., Leuthner, T.C., Di Giulio, R.T., 2017. Resistance to polycyclic aromatic hydrocarbon toxicity and associated bioenergetic consequences in a population of Fundulus heteroclitus. Ecotoxicology 26, 435–448. https://doi.org/10.1007/s10646-017-1775-6. Lowry, O.H., Rosebrough, N.J., Farr, A.L., Randall, R.J., 1951. Protein measurement with the folin phenol reagent. J. Biol. Chem. 193, 265–275. Mardirosian, M.N., Ceschin, D.G., Lascano, C.I., Venturino, A., 2017. Molecular effectors in the chronic exposure to arsenic as early and sensitive biomarkers in developing Rhinella arenarum toads. Aquat. Toxicol. 186, 19–27. https://doi.org/10.1016/j. aquatox.2017.02.019. Martínez-Gómez, C., Fernández, B., Valdés, J., Campillo, J.A., Benedicto, J., Sánchez, F., Vethaak, A.D., 2009. Evaluation of three-year monitoring with biomarkers in fish following the Prestige oil spill (N Spain). Chemosphere 74, 613–620. https://doi.org/ 10.1016/j.chemosphere.2008.10.052. Monza, L.B., Loewy, R.M., Savini, M.C., Pechen de d’Angelo, A.M., 2013. Sources and distribution of aliphatic and polyaromatic hydrocarbons in sediments from the Neuquen River, Argentine Patagonia. J. Environ. Sci. Health Part A Toxic/Hazard. Subst. Environ. Eng. 48, 370–379. https://doi.org/10.1080/10934529.2013.728097. Oberdörster, E., Brouwer, M., Hoexum-Brouwer, T., Manning, S., McLachlan, J.A., 2000. Long-term pyrene exposure of grass shrimp, Palaemonetes pugio, affects molting and reproduction of exposed males and offspring of exposed females. Environ. Health Perspect. 108, 641–646. Pacheco, M., Santos, M.A., 2001. Biotransformation, endocrine, and genetic responses of Anguilla anguilla L. to petroleum distillate products and environmentally contaminated waters. Ecotoxicol. Environ. Saf. 49, 64–75. Rewitz, K., Styrishave, B., Andersen, O., 2003. CYP330A1 and CYP4C39 enzymes in the shore crab Carcinus maenas: sequence and expression regulation by ecdysteroids and xenobiotics. Biochem. Biophys. Res. Commun. 310, 252–260. Rhee, J.S., Kim, B.M., Choi, B.S., Choi, I.Y., Wu, R.S., Nelson, D.R., Lee, J.S., 2013. Whole spectrum of cytochrome P450 genes and molecular responses to water-accommodated fractions exposure in the marine medaka. Environ. Sci. Technol. 47, 4804–4812. https://doi.org/10.1021/es400186r. Sandrini-Neto, L., Pereira, L., Martins, C.C., Silva de Assis, H.C., Camus, L., Lana, P.C., 2016. Antioxidant responses in estuarine invertebrates exposed to repeated oil spills: effects of frequency and dosage in a field manipulative experiment. Aquat. Toxicol. 177, 237–249. https://doi.org/10.1016/j.aquatox.2016.05.028. Sies, H., 1993. Strategies of antioxidant defense. Eur. J. Biochem. 215, 213–219. Song, S.B., Xu, Y., Zhou, B.S., 2006. Effects of hexachlorobenzene on antioxidant status of liver and brain of common carp (Cyprinus carpio). Chemosphere 65, 699–706. https:// doi.org/10.1016/j.chemosphere.2006.01.033. Uguz, C., Iscan, M., Ergüven, A., Isgor, B., Togan, I., 2003. The bioaccumulation of nonyphenol and its adverse effect on the liver of rainbow trout (Onchorynchus
5. Conclusions Amphipods from DS present important differences in the basal levels of antioxidant and detoxification activities compared to LB, probably related to the long-term exposure to hydrocarbons in their habitat. LB amphipods, in turn, show significant responses to short-term WAF exposure in the detoxifying and antioxidant systems. Biochemical parameters, such as CYP450 and CAT activities, could be used as oil pollution biomarkers as a useful tool in environmental monitoring and risk assessment programs. Conflicts of interest The authors have no competing interests to declare. Acknowledgements This work was supported by Grant 04A124 from the Universidad Nacional del Comahue and PICT 2013-1415 from Agencia Nacional de Promoción Científica y Tecnológica. References Achuba, F.I., Osakwe, S.A., 2003. Petroleum-induced free radical toxicity in African catfish (Clarias gariepinus). Fish Physiol. Biochem. 29, 97–103. Beers, R.F., Sizer, I.W., 1952. A spectrophotometric method for measuring the breakdown of hydrogen peroxide by catalase. J. Biol. Chem. 195, 133–140. Bhattacharyya, S., Klerks, P.L., Nyman, J.A., 2003. Toxicity to freshwater organisms from oils and oil spill chemical treatments in laboratory microcosms. Environ. Pollut. 122, 205–215. https://doi.org/10.1016/S0269-7491(02)00294-4. Borgmann, U., Grapentine, L., Norwood, W.P., Bird, G., Dixon, D.G., Lindeman, D., Chapman, P.M., Goulet, R.R., Wang, F., 2005. Sediment toxicity testing with the freshwater amphipod Hyalella azteca: relevance and application. Chemosphere 61, 1740–1745. https://doi.org/10.1016/j.chemosphere.2005.04.011. Bouvier, J.C., Boivin, T., Beslay, D., Sauphanor, B., 2002. Age-dependent response to insecticides and enzymatic variation in susceptible and resistant codling moth larvae. Arch. Insect Biochem. Physiol. 51, 55–66. Chaty, S., Rodius, F., Vasseur, P., 2004. A comparative study of the expression of CYP1A and CYP4 genes in aquatic invertebrate (freshwater mussel, Unio tumidus) and vertebrate (rainbow trout, Oncorhynchus mykiss). Aquat. Toxicol. 69, 81–94. Del Brio, J., Montagna, C.M., Lares, B.A., Parolo, M.E., Venturino, A., 2018. Chemical characterization and toxicity of water-accommodated fraction of oil on the South
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Environmental Toxicology and Pharmacology 70 (2019) 103191
J. Del Brio, et al. mykiss). Environ. Res. 92, 262–270. https://doi.org/10.1016/s0013-9351(03) 00033-1. Venturino, A., Anguiano, O.L., Gauna, L., Cocca, C., Bergoc, R.M., Pechen de D’Angelo, A.M., 2001. Thiols and polyamines in the potentiation of malathion toxicity in larval stages of the toad Bufo arenarum. Comp. Biochem. Physiol. C Toxicol. Pharmacol. 130, 191–198. Webb, D., 2011. Freshwater shrimp (Palaemonetes australis) as a potential bioindicator of crustacean health. Environ. Monit. Assess. 178, 537–544. https://doi.org/10.1007/ s10661-010-1711-1. Wheelock, C.E., Miller, J.L., Miller, M.J., Phillips, B.M., Gee, S.J., Tjeerdema, R.S., Hammock, B.D., 2005. Influence of container adsorption upon observed pyrethroid toxicity to Ceriodaphnia dubia and Hyalella azteca. Aquat. Toxicol. 74, 47–52. https:// doi.org/10.1016/j.aquatox.2005.04.007. Wu, H.Y., Xu, H.X., Hong, Y.B., Zhang, J.F., Wu, J.C., 2011. The use of biomarkers in the antioxidant responses of Daphnia magna to the acute and chronic exposure to no. 20
diesel oil and 2,4-dichlorophenol. Chem. Spec. Bioavailab. 23, 80–87. https://doi. org/10.3184/095422911X13027015402949. Yin, Y., Jia, H., Sun, Y., Yu, H., Wang, X., Wu, J., Xue, Y., 2007. Bioaccumulation and ROS generation in liver of Carassius auratus, exposed to phenanthrene. Comp. Biochem. Physiol. C Toxicol. Pharmacol. 145, 288–293. Zhang, J.F., Shen, H., Wang, X.R., Wu, J.C., Xue, Y.Q., 2004a. Effects of chronic exposure of 2,4-dichlorophenol on the antioxidant system in liver of freshwater goldfish Carassius auratus. Chemosphere 55, 167–174. Zhang, J.F., Sun, Y.Y., Shen, H., Liu, H., Wang, X.R., Wu, J.C., Xue, Y.Q., 2004b. Antioxidant response of Daphnia magna exposed to no. 20 diesel oil. Chem. Spec. Bioavailab. 16, 139–144. Zhang, J.F., Wang, X.R., Guo, H.Y., Wu, J.C., Xue, Y.Q., 2004c. Effects of water-soluble fractions of diesel oil on the antioxidant defenses of the goldfish, Carassius auratus. Ecotoxicol. Environ. Saf. 58, 110–116. https://doi.org/10.1016/j.ecoenv.2003.08. 025.
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