Effects of endosulfan, thiamethoxam, and indoxacarb in combination with atrazine on multi-biomarkers in Gammarus kischineffensis

Ecotoxicology and Environmental Safety 147 (2018) 749–758

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Effects of endosulfan, thiamethoxam, and indoxacarb in combination with atrazine on multi-biomarkers in Gammarus kischineffensis

MARK

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Özlem Demircia, , Kemal Güvenb, Dilek Asmac, Serdal Öğütd, Pelin Uğurlue a

Science Faculty, Department of Biology, Dicle University, 21280, Turkey Science Faculty, Department of Molecular Biology and Genetics, Dicle University, 21280, Turkey c Science Faculty, Department of Biology, Inonu University, 21280, Turkey d School of Health, Department of Nutrition and Dietetics, Adnan Menderes University, 09100, Turkey e Science and Technology Application and Research Center, Dicle University, 21280, Turkey b

A R T I C L E I N F O

A B S T R A C T

Keywords: Pesticide mixture Gammarus kischineffensis Atrazine Endosulfan Indoxacarb Thiamethoxam

Studies addressing the toxicity of pesticides towards non-target organisms focus on the median lethal concentration and biochemical response of individual pesticides. However, when determining environmental risks, it is important to test the combined effects of pesticides, such as insecticides and herbicides, which are frequently used together in agricultural areas. Here we aimed to investigate the toxic effects of the combined use of the herbicide atrazine and the insecticides, endosulfan, indoxacarb, and thiamethoxam on Gammarus kischineffensis. To do this, we tested the activities of oxidative stress, detoxification, and neurotoxicity biomarkers. Compared to atrazine alone, we detected higher glutathione-S-transferase, catalase and superoxide dismutase activities (oxidative stress biomarkers) when atrazine was combined with either endosulfan or indoxacarb. However, higher IBR values were determined in organisms where pesticide mixtures were used according to individual use. Based on these results, mixtures of atrazine and other pesticides may cause synergistic effects and may be evidence of increased toxicity and oxidative stress.

1. Introduction More than two million tons pesticides were in agriculture in 2006 and 2007, including pesticides, insecticides, and fungicides (Parrón et al., 2014). Insecticides, fungicides, and herbicides are often applied together as an admixture. Agricultural and urban activities have contributed to an increase in pesticide mixtures in freshwater and marine systems (Güngördü et al., 2016). Pesticides enter aquatic systems by various atmospheric and hydrologic processes, including direct application, industrial and urban discharges, and surface runoff from nonpoint sources (Sharma, 1990). The simultaneous existence of multiple anthropogenic chemicals in nature makes it difficult to understand how ecological communities are affected by them. Since the toxicity of the chemicals is usually made separately, the effect of the mixture is neglected (Beyer et al., 2014). But for two decades, research into the effects of various pollutant mixtures on aquatic populations has increased considerably (Hayes et al., 2006; Mofeed and Mosleh, 2013; Oruç and Üner, 2000; Yologlu and Ozmen, 2015). Atrazine (ATR) is a widely used herbicide and, thus, a frequent contaminant in surface and ground waters (Goldman, 1994). Atrazine

functions by binding to the plastoquinone binding protein in photosystem II, resulting in plant death caused by starvation and oxidative damage caused by a breakdown in the electron transport process (Yu et al., 2011). Many studies have evaluated the potential risk of ATR in rats, fish, freshwater mollusks, and other animals (Pogrmic-Majkic et al., 2012; Roses et al., 1999; Wiegand et al., 2000). However, few studies have examined potential interactions between triazine herbicides and insecticide mixtures (Jin-Clark et al., 2002). Organochlorine pesticides require special attention because of their high stability and toxicity to aquatic organisms (Ballesteros et al., 2009). Endosulfan (END) is an organochlorine insecticide that was commonly used before being banned worldwide. It is possible that END continues to cause serious damage to ecosystems, especially in aquatic environments (Dayakar et al., 2015; Shao et al., 2012; Singh and Singh, 2017). Because organochlorine pesticides are hydrophobic, biologically stable and can dissolve in oil, they can easily accumulate in the adipose tissue of organisms. Most of the organochlorine pesticides are highly durable (Darko et al., 2008). Indoxacarb (IND) is the first commercialized pyrazoline type insecticide. IND has high insecticidal activity and low toxicity to non-target organisms. Because indoxacarb is a pro-oxidant, it

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Corresponding author. E-mail addresses: [email protected] (Ö. Demirci), [email protected] (K. Güven), [email protected] (D. Asma), [email protected] (S. Öğüt), [email protected] (P. Uğurlu). http://dx.doi.org/10.1016/j.ecoenv.2017.09.038 Received 16 June 2017; Received in revised form 12 September 2017; Accepted 14 September 2017 0147-6513/ © 2017 Elsevier Inc. All rights reserved.

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maintained in glass aquariums containing a 50:50 mixture of fieldsampled freshwater and dechlorinated tap water. The animals were maintained in a special room at 18 ± 1 °C under an artificial light regime (13 h light: 11 h dark). Dried willow tree (Salix sp., Salicaceae) leaves were collected from the natural environment of gammarids, and the leaves were conditioned in field-sampled water for at least two weeks before use (Cold and Forbes, 2004), after which individuals were fed the decomposed leaves during an adaptation period (Gerhardt, 2011).

needs to be metabolized to the decarboxylated form to become biologically active. Its metabolite is known to inhibit sodium channels in cultured insect neurons (Jones and Bryant, 2012). However, little is known about the effects of these compounds on aquatic organisms (Zhao et al., 2003). The synthetic organic pesticide thiamethoxam (THI) is a broad spectrum neonicotinoid insecticide. Neonicotinoids act on nicotinic acetylcholine receptors (Matsuda et al., 2001). Xenobiotics generate an important source of reactive oxygen species (ROS). ROS are released by various metabolic reactions, such as biosynthesis and biodegradation, biotransformation of xenobiotics, cellular respiration, and phagocytosis. Exposure to various environmental toxicants discharged from agriculture, industry, tobacco smoke, or pollution accidents likely increase ROS production. Excessive ROS production causes damage to cellular components, leading to a process known as oxidative stress (Ojha et al., 2011). Oxygen free radical enzymatic scavengers, such as superoxide dismutase (SOD), catalase (CAT), glutathione-S-transferase (GST), glutathione peroxidase (GPx) and glutathione reductase (GR) provide protection from the deleterious effect of ROS (Banerjee et al., 1999). Although aquatic invertebrates have been employed as biological indicators, the use of acetylcholinesterase (AChE) inhibition as a biomarker has been largely neglected in these species (Hynea and Maherb, 2003; Vioque-Fernández et al., 2007). Many insecticides are toxic to animals because they irreversibly inhibit AChE, an enzyme of the animal nervous system. Thus, AChE inhibition can be used as an indicator of exposure to organophosphorus pesticides (Hynea and Maherb, 2003). Biochemical markers can provide information about the health status of organisms and, therefore, can be used as early warning signals of general or particular stress (Korte et al., 2000). Conventional water quality monitoring studies typically determine the optimum threshold of one chemical. However, these chemistries have similar or different modes of action, it is difficult to predict their mixing effects (Beyer et al., 2014). Therefore, studies investigating the combined effect are important to understanding the real biological effect and developing environmental protection strategies (Donner et al., 2010; Langston et al., 2007). Because chemical toxicants can generate distinct measurable biological responses in affected organisms, biomarkers can be used to assess toxicant-induced changes at ecological and biological levels (Adams and Rowland, 2003). In this study, G. kischineffensis, a gammaridean arthropod, was used as the test organism. Gammarus spp is quite common in freshwater and is a very dominant part of the macroinvertebrates in the benthic zone. Gammarids are both herbivorous or carnivorous feeding on small invertebrates and carvings (MacNeil et al., 1997). Gammaridean species are important bioindicators for both marine and freshwater aquatic environments (Gerhardt et al., 2011; Woods et al., 2002). Moreover, because of restrictions on studies using mammalians, invertebrate studies have recently gained greater attention (Kandárová and Letašiová, 2011). The combined effect of pesticides has been shown in many studies (Bacchetta et al., 2014; Fatima et al., 2007; Mofeed and Mosleh, 2013). However, these studies have shown that the effects of chemical mixtures on different organisms can be quite different. Therefore, the main objective of this study is to determine the toxic effects of sublethal concentrations of four pesticides (ATR, END, IND and THI) on G. kischineffensis, an important bioindicator, both alone and as part of a mixture.

2.2. Acute toxicity experiment Before experiments, all glass materials used in the experiments were prepared according to the methods book of the American Public Health Association (Federation and Association, 2005). Organisms were exposed to pesticides in well-aerated 2 L glass jars that contain 1 L of water. During 96 h, 20 individuals in each of the triplicate were used for each concentration tested. In all experiments carried out in the aquaria, half of the total water volume was replenished at 24 h intervals, adding pesticide to maintain fixed concentrations. After range finding tests, eleven concentrations of ART (1–50 mg/L), twelve concentrations of END (1–4 mg/L), eleven concentrations of IND (15–65 mg/L), and thirteen concentrations of THI (1–30 mg/L) were selected to determine their LC50 values. Analytical standards of all pesticides were used throughout the study. The purity ratings of ATR, END, IND, and THI used were ≥ 98.0,%, 71.0%, ≥ 95.0%, and ≥ 98.0%, respectively. 2.3. Combined effect After we had determined the 96 h LC50 values of the four tested pesticides, combinations of the 1/100 of the LC50 concentration of the herbicide ATR with the same concentration of each insecticide, THI, IND, and END were tested separately to detect the combined effects of pesticides. Also, the 96 h LC50/100 concentrations of these four pesticides were applied singly and used for comparison. All experiments were conducted for 96 h. The organisms were fed with decomposed willow (Salix sp.) leaves collected from the sampling site. All data obtained are the mean of three experiments. 2.4. Enzyme activity assays After sublethal exposure, frozen animal tissues were homogenized in sodium phosphate buffer (0.1 M, pH 6.5) and centrifuged at 25,000g for 10 min at 4 °C. The supernatant was then transferred into clean microfuge tubes. GR, SOD, GST and AChE activities were determined spectrophotometrically at appropriate wavelengths using a microplate reader (VersaMax, Molecular Devices Corp., USA) at 25 °C. CAT activity was determined using a SHIMADZU UV–Visible Spectrophotometer (UV1601). CAT activity was determined by measuring absorbance decrease at 240 nm due to hydrogen peroxide decomposition, according to the method of Luck (1963). One unit of enzyme activity was defined as the amount of the enzyme decreasing 1 μM H2O2 per min. GR activity was measured using the method described modified by (Cribb et al., 1989). The reaction solution included 0.1 mM 5,5′-Dithiobis(2-nitrobenzoic acid) (DTNB) as substrate, 1.2 mM NADPH and 10 µL of homogenate supernatant in a total volume of 190 µL. After the reaction initiation, 20 µL of 3.25 mM Glutathione oxidized (GSSG) was added; Glutathione reduced (GSH) was generated from GSSG. DTNB reduction was monitored at 405 nm, and the extinction coefficient for DTNB (ε = 14,151 M−1 cm−1) was used to calculate the enzyme activity. GST activity was determined using the method described by Habig et al. (1974). To measure the activity of GST, the reaction solution consisted of 1-chloro, 2-4dinitrobenzene (CDNB) as a substrate and reduced glutathione (GSH) as a cofactor. The reaction solution

2. Materials and methods 2.1. Collection and maintenance of animals G. kischineffensis samples were collected from the Tigris River using laboratory test sieves. Samples were collected from the deepest part of a stream under stones and leaves (36ᵒ 54′ 56.84′′ N, 40ᵒ 16′ 28.50′′ E). Only healthy adult individuals, which have pre-copulation behavior, were selected for the experiments. G. kischineffensis samples were 750

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contained 0.1 M potassium phosphate buffer (pH 6.5), 1 mM GSH, 1 mM CDNB and 10 µL sample volume. Enzyme activity was determined by monitoring changes in absorbance at 340 nm, which related to the rate of CDNB conjugation with GSH. The GST activity is expressed as μmol/min/mg protein. The AChE activity was measured following the methods by Ellman et al. (1961) and that modified for microplate readers by Ozmen et al. (2008). The reaction solution was prepared to the final concentration of 0.7 mM acetylthiocholine iodide (ACTI) and 0.14 mM DTNB in Trizma buffer (0.1 M, pH 8.0). Enzyme kinetics were monitored at 412 nm for 3 min. A commercial SOD assay kit (Sigma-Aldrich) was used to determine SOD activity at 450 nm wavelengths. Total protein concentrations of each sample were measured according to Bradford method using bovine serum albumin as a standard (Bradford, 1976), employing a Bio-Rad protein assay kit.

Table 1 The 96 h LC50 value of ATR, END, IND, and THI on G. kischineffensis (ATR: Atrazine; END: Endosulfan; IND: Indoxacarb and THI: Thiamethoxam). Test range (AI)

ATR (mg/L) END (µg/L) IND (mg/L) THI (mg/L)

1.0–50.0 1.0–4.0 15.0–65.0 1.0–30.0

No tested cons.

10 10 10 10

LC50 24 h

48 h

72 h

96 h

35.5 3.9 54.4 25.5

26.8 2.4 45.2 19.4

23.4 1.8 44.9 15.1

18.9 1.5 39.0 8.9

10 individual organisms were used for each treatment. Values in parentheses for LC50 values represent 95% confidence limits.

of active ingredient (AI) L−1, respectively. The 96 h LC50 value of END is much lower than the other pesticides. The 24, 48, 72 and 96 h LC50 values of the pesticides are shown in Table 1.

2.5. Calculations for combined effects Living organisms are exposed to the combined effects of pesticides at concentrations derived from LC50/100. The percentages of the specific enzyme activities obtained from the pesticide exposures were calculated by dividing these values by that of the control group and then multiplying by 100. Exposed groups specific activity × 100 Combined Effect% = Control specific activity

3.2. Time-dependent changes in biomarker responses in G. kischineffensis Here, statistical analysis was carried out for single treatments, as well as combined treatments of all pesticides, compared to their untreated controls, as well as the combined treatments compared with individual treatments. We found that the enzyme activities for single and combined treatments differed depending on exposure times (Tables 2–4). GST activity in the ATR-treated groups was different from the control for all exposure times (p < 0.05). The GST activity of ATR +END was higher than the ATR and END individual treatments, approximately 1.5-fold and 1.7-fold (72 h); 1.3-fold and 1.5-fold (96 h), respectively. Similarly, GST activity of ATR+IND was higher than the single use of ATR and IND, approximately 1.4-fold and 1.8-fold (72 h); 1.4-fold and 1.2-fold (96 h), respectively. (However, the GST activity after 24 h exposure to the ATR+THI mixture was lower than the single treatment groups for the same exposure period (p < 0.05). We found that the CAT activity, after the combined exposures for 72 h, was higher than the single ATR and END treatment groups, approximately 13.6-fold and 3.5-fold respectively (p < 0.05). On the other hand, comparison of the single IND treatment group with combined effect with ATR showed that the combined effects of ATR increased the CAT activity compared to single IND application at 24 and 48 h, by approximately 2.1-fold and 1.5-fold (ATR); 5.0-fold and 1.2fold (IND), respectively (p < 0.05). We found that, after 24 and 96 h of exposure, the SOD activity was greater in the ATR+END group than the single treatments, by approximately 1.5-fold and 1.2-fold (ATR); 2.0-fold and 1.6-fold (END), respectively (p < 0.05). Also, after 48 and 72 h of exposure, the SOD activity was greater in the ATR+IND group than the single treatments, by approximately 1.2-fold and 1.2-fold (ATR); 1.3-fold and 1.2-fold (IND), respectively (p < 0.05). Compared to the single treatments, SOD activity was greater in the ATR+THI group after 96 h of exposure 1.1fold (single ATR) and 1.5-fold (single THI) (p < 0.05). Compared to the single treatments, the GR activity was reduced in the combination treatments at 24 h, approximately 1.7-fold (END); 1.4fold (IND) and 1.5-fold (single THI) (p < 0.05). ATR+IND had higher GR activity than the single treatments at 72 h of exposures, approximately 1.5-fold (ATR); and 0.8-fold (IND) (p < 0.05). In the ATR+END group, AChE activity was reduced at 48 h compared to the single exposures, approximately 1.6-fold (ATR) and 1.6fold (END) (p < 0.05). Compared with single use of pesticides, ATR +IND caused suppression of AChE activity at 24 h and 48 h, approximately 2.9-fold and 1.9-fold (ATR); 1.8-fold and 1.2-fold (IND), respectively (p < 0.05) And ATR+THI caused suppression of AChE activity compared with single use of ATR and THI at only 24 h approximately 3.0-fold (ATR) and 1.2-fold (END) (p < 0.05).

2.6. Statistical analysis Probit analysis was used determine the 24, 48, 72 and 96 h LC50 values (SPSS v. 15.0) (Finney, 1952). The statistical package program SPSS version 15.0 software was used to evaluate the enzyme activity assay data. Kruskal-Wallis test was used to examine the differences between enzyme activities at the different time periods and, thus, to test the time-dependent change in each dose. After the Kruskal-Wallis test, the Mann-Whitney U-test was used for statistical determination of pairwise comparisons. In these tests, the level of significance was set at p < 0.05. 2.7. Assessment of integrated biomarker response The integrated biomarker response (IBR) was calculated by combining all assayed biomarker responses into one general stress index. This allowed us to evaluate the potential risk of pesticides, either singularly or in combination. The IBR index was calculated as follows: for each biochemical marker, mean (m) and standard deviation (SD) were estimated for each exposure (Arzate-Cárdenas and Martínez-Jerónimo, 2012). The mean value for each response was separately evaluated and standardized using the formula Y = (X-m)/SD, where Y is the standardized value of the biochemical marker, X is the mean value of a biochemical marker for each treatment, and m is the mean of the biochemical markers calculated for all treatments. Z values were computed as Z = Y or Z = − Y, depending on the biological effect (inhibition or stimulation, respectively). Afterward, score (S) was assessed by the formula S = |min|+Z, where |min| is the absolute value of the minimum of all treatment groups for every biochemical marker. For estimation of the IBR index, these scores were utilized: [(S1 × S2)/2 + (S2 × S3)/2 +…(Sn − 1 × Sn)/2]. The values estimated were divided by the number of biochemical markers calculated to yield a normalized IBR (Broeg and Lehtonen, 2006). 3. Results 3.1. Acute toxicity tests We calculated the 96 h LC50 values for ATR, END, IND, and THI against G. kischineffensis as 18.96 mg, 1.56 µg, 39.04 mg, and 8.985 mg 751

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Table 2 The percentage activities of selected enzymes in G.kischineffensis after 96 h single and combine ATR and END exposure. (ATR: Atrazine and END: Endosulfan).

ATR

END

ATR+END

Exposure period (h)

GST

CAT

SOD

GR

AChE

24 48 72 96 24 48 72 96 24 48 72 96

121.9 ± 2.8a 150.6 ± 4.0a 129.6 ± 6.0a 128.8 ± 3.8a 58.5 ± 4.4a 110.1 ± 1.1 108.5 ± 1.8 106.4 ± 3.8 113.4 ± 11.5b 133.7 ± 7.4b,c 187.6 ± 8.5a,b,c 165.3 ± 12.1b,c

88.5 ± 0.7a 89.4 ± 4.6 46.0 ± 4.3a 95.3 ± 21.7 113.0 ± 6.8 182.6 ± 6.9a 176.7 ± 21.7a 391.5 ± 19.3a 122.5 ± 15.9a,b 203.7 ± 19.6b 626.2 ± 23.9b,c 405.7 ± 33.1a,b

123.7 ± 7.3a 145.8 ± 8.4a 142.5 ± 6.2a 133.0 ± 3.9a 95.4 ± 2.9 96.1 ± 6.4 88.3 ± 8.1a 98.0 ± 2.5 190.8 ± 7.3a,b,c 130.1 ± 13.6c 136.8 ± 1.4c 159.5 ± 15.9b,c

192.5 ± 6.8a 97.4 ± 5.0 78.7 ± 8.9a 126.8 ± 3.9a 139.2 ± 4.4a 109.3 ± 0.9 145.2 ± 1.6a 79.7 ± 7.8a 78.9 ± 12.2b,c 98.0 ± 8.5b 92.6 ± 13.1 120.7 ± 9.1c

244.7 ± 11.5a 191.4 ± 11. 4a 172.9 ± 9.4a 168.4 ± 10.1a 146.9 ± 4.0a 186.1 ± 13.3a 122.4 ± 9.9a 118.9 ± 16.5a 120.6 ± 10.8b 115.8 ± 3.4b,c 118.9 ± 7.2c 114.3 ± 7.2c

Percentage enzyme activities were expressed as nmol min−1 mg protein −1 ± standard error. Organisms were treated with 0.18 mg/L ATR and 0.05 µg L−1 END. a Statistically significant when compared with control (p < 0.05). b Statistically significant when compared with the ATR exposure group (p < 0.05). c Statistically significant when compared with the END exposure group (p < 0.05).

When used in combination, pesticides might produce unexpected toxicity responses and more toxic levels than expected. Recently, awareness about the investigation of the effect of mixtures of anthropogenic pollutants on the aquatic system has increased considerably (Feng et al., 2015). The main goal of the present work was to evaluate possible synergistic, additive and antagonistic effects of pesticides on G. kichneffensis, using multi-biochemical biomarkers. The inhibition or induction of biomarkers is useful as an environmental tool to assess the exposure and potential effects of xenobiotics on organisms (Rendón-von Osten et al., 2005). Organisms are exposed to concentrations of xenobiotics in their natural environment; often well below the LC50 values determined in laboratories. Therefore, after finding the LC50 values of the pesticides, we tried to determine the combined effect using 1/100 of the LC50 value for each pesticide. Previous studies have shown that the concentration of applied pesticide is not always the same as the concentration that is being exposed (Mofeed and Mosleh, 2013). However, ATR, END, IND, and THI have relatively long half-lives, approximately 90, 92, 38 and 250 days, respectively (determined at pH 7) (Jayaprabha and Suresh, 2016; Maienfisch et al., 2001; Moncada, 2003; Schottler et al., 1994). In our study, carried out in the aquaria, half of the total water volume was replenished at 24 h intervals, adding pesticide to maintain fixed concentrations. Oxidative stress is induced by pesticides and pesticide mixtures. ATR is known to increase the toxicity of many insecticides when

3.3. Integrated biomarker response (IBR) Fig. 1. shows the IBR indexes of all biomarker values for the ATR and END exposures, both singularly and in combination. We found that IBR is increased at 72 and 96 h of exposure in ATR+END compared to the single treatments. In contrast, as seen in Fig. 2. IND singularly caused a dramatic increase in IBR levels at 96 h. Whereas IND+ATR caused an increase in IBR levels only at 48 and 72 h, compared to the single treatments. Fig. 3. shows that only the single THI exposure increased the IBR index at 72 h.

4. Discussion Atrazine is the most popular triazine herbicide in the world. And atrazine is currently used in over 60 countries and most of them are developing countries (Pathak and Dikshit, 2012; Roustan et al., 2014). Some studies have been carried out on the single or combined effects of atrazine on aquatic organisms such as fish, clams, amphibians, planktons. (dos Santos and Martinez, 2014; Fatima et al., 2007; Hayes et al., 2006; Hoagland et al., 1993; Mehler et al., 2008). Perhaps the most serious problem among these xenobiotics is pesticides because they are specifically designed to kill an organism (Hanazato, 2001). Also, pesticides are often overused on agricultural land, as a mixture rather than individually. Ecotoxicological evaluations have led to the underestimation of toxicity when based on individual effects (Gilliom, 2007).

Table 3 The percentage activities of selected enzymes in G. kischineffensis after 96 h single and combine ATR and IND exposure (ATR: Atrazine and IND: Indoxacarb). Exposure period (h) ATR

IND

ATR+IND

24 48 72 96 24 48 72 96 24 48 72 96

GST

CAT a

121.9 ± 2.8 150.6 ± 4.0a 129.6 ± 6.0a 128.8 ± 3.8a 84.8 ± 6.3a 116.3 ± 12.2a 102.4 ± 2.8 164.3 ± 1.9a 107.2 ± 3.7b,c 145.0 ± 3.8a,b 186.7 ± 0.9a,b,c 179.4 ± 6.7b,c

SOD a

88.5 ± 0.7 89.4 ± 4.6 46.0 ± 4.3a 95.3 ± 21.7 37.1 ± 2.4a 115.5 ± 4.1 362.9 ± 19.3a 655.3 ± 8.3a 186.910.1b,c 139.1 ± 14.0b,c 152.6 ± 5.0b,c 113.2 ± 6.3c

Percentage enzyme activities were expressed as nmol min−1mg protein −1 ± standard error. Organisms were treated with 0.18 mg/L ATR and 0.39 mg/L IND. a Statistically significant when compared with control (p < 0.05). b Statistically significant when compared with the ATR exposure group (p < 0.05). c Statistically significant when compared with the IND exposure group (p < 0.05).

752

GR a

123.7 ± 7.3 145.8 ± 8.4a 142.5 ± 6.2a 133.0 ± 3.9a 95.7 ± 5.9 127.4 ± 3.9a 136.1 ± 5.9a 162.7 ± 2.9a 123.6 ± 4.1c 167.1 ± 5.6b,c 163.0 ± 7.4b,c 145.5 ± 1.2b,c

AChE a

192.5 ± 6.8 97.4 ± 5.0 78.7 ± 8.9a 126.8 ± 3.9a 114.9 ± 7.2 117.0 ± 10.0 140.3 ± 3.7a 236.5 ± 4.1a 85.1 ± 3.2b,c 116.3 ± 4.1b 163.2 ± 1.3b,c 129.9 ± 10.3c

244.7 ± 11.5a 191.4 ± 11. 4a 172.9 ± 9.4a 168.4 ± 10.1a 147.1 ± 6.3 124.1 ± 6.9 159.5 ± 4.7 146.2 ± 7.6 81.7 ± 1.9a,b,c 102.9 ± 3.1b,c 179.1 ± 7.4a,b 149.2 ± 1.8a,b

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Table 4 The percentage activities of selected enzymes in G. kischineffensis after 96 h single and combine ATR and THI exposure (ATR: Atrazine and THI: Thiamethoxam).

ATR

THI

ATR+THI

Exposure period (h)

GST

CAT

SOD

GR

AChE

24 48 72 96 24 48 72 96 24 48 72 96

121.9 ± 2.8a 150.6 ± 4.0a 129.6 ± 6.0a 128.8 ± 3.8a 129.8 ± 1.2a 109.5 ± 7.4 114.8 ± 2.9a 99.9 ± 8.2 105.9 ± 9.8b 153.9 ± 5.6c 130.8 ± 7.7c 131.9 ± 3.6a,c

88.5 ± 0.7a 89.4 ± 4.6 46.0 ± 4.3a 95.3 ± 21.7 245.7 ± 19.6a 181.7 ± 14.9a 303.1 ± 4.9a 155.6 ± 4.2a 98.1 ± 7.3c 97.2 ± 7.5c 73.9 ± 3.4b,c 80.0 ± 4.6b,c

123.7 ± 7.3a 145.8 ± 8.4a 142.5 ± 6.2a 133.0 ± 3.9a 78.5 ± 3.4 120.3 ± 3.5 132.4 ± 5.7a 97.5 ± 3.1 100.2 ± 7.5b,c 138.8 ± 6.9c 145.0 ± 4.5 146.8 ± 0.9a,b,c

192.5 ± 6.8a 97.4 ± 5.0 78.7 ± 8.9a 126.8 ± 3.9a 114.9 ± 3.7 131.5 ± 6.0a 138.4 ± 7.1a 123.7 ± 4.9a 76.7 ± 6.6b,c 138.5 ± 6.4b 154.5 ± 11.0b 145.0 ± 14.6

244.7 ± 11.5a 191.4 ± 11. 4a 172.9 ± 9.4a 168.4 ± 10.1a 99.3 ± 5.9 102.2 ± 7.6 114.9 ± 8.7a 105.2 ± 3.3a 80.9 ± 2.5b,c 145.7 ± 8.8b,c 140.8 ± 6.8c 139.1 ± 6.4b,c

Percentage enzyme activities were expressed as nmol min−1 mg protein −1 ± standard error. Organisms were treated with 0.18 mg/L ATR and 0.08 mg/L THI. a Statistically significant when compared with control (p < 0.05). b Statistically significant when compared with the ATR exposure group (p < 0.05). c Statistically significant when compared with the THI exposure group (p < 0.05).

classes, oxidative stress and toxicity may occur at different levels. In our study, compared to the single pesticide exposures, the combination of END with the herbicide ATR increased CAT activity after 72 h of exposure, and CAT activity was increased after 24 and 48 h of exposure in the IND and ATR mixture. However, CAT activity in the THI and ATR mixture exposed animals was reduced after all exposure periods (Tables 2–4) (Figs. 4–6) (p < 0.05). These findings suggest that there is heterogeneity in the enzyme responses within the same organism and indicate that different pesticides affect ROS production differently. The present study also shows that all tested insecticides, in combination with ATR, induced SOD activity at some time intervals, varying between 24 and 96 h. These results suggest that the mixture of pesticides leads to the generation of superoxide anions and therefore hydrogen peroxide (H2O2), induces the activation of the enzymatic antioxidant defense system for preventing oxidative damage (Bacchetta et al., 2014). Oruç et al. (2004) studied the single or combined effects of the pesticides 2,4-D, and azinphosmethyl on SOD activity in two fish species (Oreochromis niloticus and Cyprinus carpio). They reported that a significant change was not observed in SOD enzyme activity in the liver and brain tissues in the fish species O. niloticus and C. carpio, compared to the control. However, a significant induction of SOD activity in the gills was obtained when the pesticides were applied both separately and

applied in combination. For instance, it was shown that ATR enhances parathion (an organophosphate insecticide) toxicity in the larvae of mosquito (Aedes aegypti), and Drosophila melanogaster (fruit fly), as well as increases carbofuran (carbamate insecticide) toxicity in Musca domestica (housefly) (Anderson and Lydy, 2002; Lichtenstein et al., 1979, 1974). Moreover, the toxicity of OP insecticides was elevated when applied in combination with ATR in C. tentans (Belden and Lydy, 2000). However, other studies indicate that ATR decreases the toxicity of methoxychlor (organochlorine insecticide) and mevinphos (phosphate insecticide) by altering the toxic action of insecticides in the midge Chironomus tentans (Pape-Lindstrom and Lydy, 1997). Antioxidant enzymes we use to determine the presence and amount of ROS formed as the result of oxidative stress caused by xenobiotics very important biomarkers (Demirci and Hamamcı, 2013). SOD is an antioxidant enzyme that catalyzes the conversion of superoxide radical (O2•-), hydrogen peroxide (H2O2) and molecular oxygen (O2). Any increase in SOD activity also causes changes in CAT activity as it also increases the production of H2O2, the SOD product. If SOD activity exceeds the activity of H2O2 scavengers, it leads to increases in radical toxicity due to H2O2 accumulation. Hydrogen peroxide (H2O2) formed in the cell is removed by CAT to prevent the formation of the hydroxyl radical (•OH) (Weydert and Cullen, 2010). Since the insecticides which we have studied in combination with atrazine belong to different

Fig. 1. IBR analysis of selected biochemical markers; GST, CAT, SOD, GR, AChE, on G. kischineffensis exposed to atrazine combination with endosulfan.

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Fig. 2. IBR analysis of selected biochemical markers; GST, CAT, SOD, GR, AChE, on G. kischineffensis exposed to atrazine combination with indoxacarb.

Güngördü et al. (2016), GR enzyme activity of three different types of amphibia exposed to a mixture of glyphosate-based and methidationbased pesticides was increased at a significant level whereas in another study GR enzyme inhibition was found in the fish (Piaractus mesopotamicus) exposed to endosulfan and amphora-cyhalothrin mixtures (Bacchetta et al., 2014). These results suggest that, in general, pesticide mixtures cause more oxidative stress toxicity than single use due to additive or synergistic interactions. It is well known that environmental contaminants detoxified by GSTs include polyaromatic hydrocarbons, pesticides, and reactive intermediates generated by phase I biotransformation and other biochemical reactions (Richardson et al., 2009; Trute et al., 2007; Xing et al., 2010, 2012). GSTs function in both antioxidant and detoxification processes, catalyzing the conjugation of various electrophilic compounds with glutathione and convert these compounds to watersoluble compounds. Bacchetta et al. (2014) observed that GST enzyme activity was induced in the liver of fresh water fish (Piaractus mesopotamicus) exposed to a combined mixture of END and lambda-cyhalothrin. Mofeed and Mosleh (2013) studied the toxic responses and antioxidative enzyme activities in the algae Scenedesmus obliquus and found that GST activity increased in the algae exposed to fenhexamid and ATR compared to single treatments (for lower doses). We also

together. Moreover, a meaningful increase was detected in CAT activity only in the liver tissues of both fish species. These results show that a xenobiotic can cause dissimilar oxidative stress in different tissues of the same organism. In addition, these results may indicate that the pesticide mixture induces synergistic interactions leading to more oxidative stresses, and thus induces CAT and SOD enzymes. GR is an enzyme that indirectly acts as an antioxidant by converting oxidized glutathione (GSSG) into reduced glutathione (GSH), which is formed during reactions catalyzed by glutathione peroxidase (GPx) and glutathione S-transferase (GST) (Pai and Schulz, 1983; Van der Oost et al., 2003). While GPx plays an important role in removing oxidative stress, GST uses GSH for the detoxification of pesticides (Banerjee et al., 1999; Chen et al., 2015). In this study, GR activity in G. kischineffensis was especially decreased after 24 h of exposure to pesticide combinations, but there was an increase for the same enzyme after 72 h exposure to combinations of ATR with IND and THI, compared to single exposures (Tables 2–4) (Figs. 4–6) (p < 0.05). When the responses to aquatic organisms exposed to different combinations of pesticides are examined in the literature, different results are presented. Similarly, in another study, after 48 h the fish that were exposed to a mixture of 2,4D and azinphos-methyl showed a significant increase in GR activity compared to the individual treatments (Oruc et al., 2004). In a study by

Fig. 3. IBR analysis of selected biochemical markers; GST, CAT, SOD, GR, AChE, on G. kischineffensis exposed to atrazine combination with thiamethoxam.

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Fig. 4. The percentage activities of selected enzymes in G. kischineffensis after 96 h single and combine ATR and END exposure.

and 0.25 mg/L concentration of chlorpyrifos on AChE enzyme in aquatic dipteran Chironomus tentans. These authors reported that there was no AChE enzyme inhibition in the single application groups of ATR and chlorpyrifos, while there was an inhibition in combined mixture exposure groups. These studies, which were in agreement with the present study, show that there is less neurotoxic effect in single use compared to combined use. Chlorpyrifos is an organophosphate insecticide, and AChE enzyme is expected to be the first target for organophosphates. However, organophosphate insecticides are supposed to be transformed into oxo analogs with oxidative activation provided by the P450 enzyme. Thus, they can only affect (inhibit) the activity of the AChE enzyme via this pathway. Most likely, ATR induces expression of the cytochrome P450 enzyme, which transforms chlorpyrifos to chlorpyrifos-oxone. This might explain why AChE is inhibited in organisms exposed to combinations of pesticides (Cremlyn, 1991). We found that levels of the CAT, SOD and GR enzymes in the animals exposed to END, IND, THI in combination with ATR were not significantly different from the control groups. However, we did detect differences in enzyme activities when comparing single and combined exposures. Exposure duration also had an effect on some biomarkers. Thus, combinations of pesticides can alter their toxicity. When evaluating the responses of multiple biomarkers, it is difficult to determine the resulting combined effect correctly. For this reason, we used IBR

showed that GST response is increased in G. kichneffensis exposed to ATR, END and IND mixtures (Tables 2–4) (Figs. 4–6) (p < 0.05). These results may indicate that the pesticide mixture induces synergistic interactions leading to greater toxicity and thus induction of the GST enzyme. Another biomarker that we use in our work is AChE, which is a very important enzyme in the nervous system. This enzyme terminates nerve impulses by providing hydrolysis of the neurotransmitter acetylcholine (Lionetto et al., 2013). The inhibition of AChE by xenobiotics causes acetylcholine accumulation and hyperstimulation of receptors. For this reason, the AChE enzyme is one of the vital biomarkers used to demonstrate the effects of environmental pollutants (Colovic et al., 2013). In our study, it was observed that combined use of ATR with END, IND and THI inhibited AChE activity especially in the first two days of treatment (Tables 2–4) (Figs. 4–6) (p < 0.05). These results may indicate that the pesticide mixture is causing more neurotoxic effects in the early hours of treatment (Lionetto et al., 2013). Xing et al. (2010) examined AChE enzyme response in the brain and muscle tissues of carps exposed to separate concentrations and a combined mixture of ATR and chlorpyrifos for 40 days. It was found that in both applications the enzyme activity was inhibited by increasing concentrations. JinClark et al. (2002) investigated the combined effect of 1, 10, 100 and 1.000 mg/L concentrations of two herbicides (atrazine and cyanazine)

Fig. 5. The percentage activities of selected enzymes in G. kischineffensis after 96 h single and combine ATR and IND exposure.

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Fig. 6. The percentage activities of selected enzymes in G. kischineffensis after 96 h single and combine ATR and THI exposure.

other pesticides. According to our IBR analysis, although IND and THI are new generation insecticides, END and ATR, which are organochlorine insecticides, produced a higher toxic response. For this reason, we conclude that the multi-biomarkers we use are suitable for determining such toxicity. However, the IBR level produced by the END and ATR mixture is high compared to the single treatments, whereas, with combinations of IND and THI with ATR, the opposite trend is observed. Thus, the combined toxicity of xenobiotics likely applied together (e.g., insecticides and herbicides) should be carefully evaluated.

analysis to avoid confusion of pesticide combination responses and to make accurate toxicological evaluations. A variety of biochemical markers is necessary to relate any treatment to response and, thus, to provide better predictive tools for quantitative monitoring of the toxicological effects of most toxicants (Venturino et al., 2003). IBR analysis can be used to determine the effects of toxicants on various species under laboratory conditions (Arzate-Cárdenas and Martínez-Jerónimo, 2012; Qu et al., 2013). An IBR index has been used to demonstrate the effects of various chemicals on amphibians (Güngördü et al., 2016; Ozmen et al., 2015). In our study, we have found that combinations of END, IND, and THI with ATR lead to higher IBR values at different exposure time compared to the single pesticide exposures (Figs. 1–3). The IBR values obtained from organisms exposed to the ATR + END mixture can be said to exhibit a significant synergistic effect, especially at 72 and 96 h, when compared with single use. However, when IBR values obtained from organisms exposed to the ATR + IND mixture are compared with single use, it can be said that there is a significant antagonistic effect after 96 h, especially when the synergistic effect is observed after 72 h. Combined treatment of ATR+THI for 48 and 96 h caused a slightly synergistic effect, while combined treatment for 24 and 72 h had an antagonistic effect on the IBR values. The results obtained outside of this study showed that the combination of END and IND with the ATR mixture resulted in more synergistic effect when compared to the THI + ATR mixture. For this reason, it can be said that the mixture of END and IND with ATR causes more stress and therefore toxic effect on G. kischineffensis. These observations were consistent with the findings of other studies using an IBR approach in mussels (Brooks et al., 2015), earthworms (Stepić et al., 2013) and fish (Kim et al., 2010; Li et al., 2011; Suman et al., 2015). Thus, IBR is useful for making integrated interpretations of the biological effects of contaminants. Our findings, as well as those of other studies, suggest that combinations of xenobiotics can induce unexpectedly severe stress responses in non-target organisms. Therefore, it is important to carefully identify the combined effects of xenobiotics commonly used in natural sites.

Acknowledgments This study is a part of a comprehensive project that is supported by The Scientific and Technical Research Council of Turkey (TUBITAK) (Project no. 111T661) and Dicle University Academic Research Project Unit (DUBAP) (Project no. 12-FF-87). We are grateful to Dr. Abbas Güngördü (Inonu Universty) for their critical evaluation of the manuscript. References Adams, W.J., Rowland, C.D., 2003. Aquatic Toxicology Test Methods. A CRC Press Company, Washington, D.C. Anderson, T.D., Lydy, M.J., 2002. Increased toxicity to invertebrates associated with a mixture of atrazine and organophosphate insecticides. Environ. Toxicol. Chem. 21, 1507–1514. http://dx.doi.org/10.1002/etc.5620210724. Arzate-Cárdenas, M.A., Martínez-Jerónimo, F., 2012. Energy reserve modification in different age groups of Daphnia schoedleri (Anomopoda: daphniidae) exposed to hexavalent chromium. Environ. Toxicol. Pharmacol. 34, 106–116. http://dx.doi.org/ 10.1016/j.etap.2012.03.003. Bacchetta, C., et al., 2014. Combined toxicological effects of pesticides: a fish multibiomarker approach. Ecol. Indic. 36, 532–538. http://dx.doi.org/10.1016/j.ecolind. 2013.09.016. Ballesteros, M.L., et al., 2009. Oxidative stress responses in different organs of Jenynsia multidentata exposed to endosulfan. Ecotoxicol. Environ. Saf. 72, 199–205. http://dx. doi.org/10.1016/j.ecoenv.2008.01.008. Banerjee, B.D., et al., 1999. Biochemical effects of some pesticides on lipid peroxidation and free-radical scavengers. Toxicol. Lett. 107, 33–47. http://dx.doi.org/10.1016/ S0378-4274(99)00029-6. Belden, J.B., Lydy, M.J., 2000. Impact of atrazine on organophosphate insecticide toxicity. Environ. Toxicol. Chem. 19, 2266–2274. http://dx.doi.org/10.1002/etc. 5620190917. Beyer, J., et al., 2014. Environmental risk assessment of combined effects in aquatic ecotoxicology: a discussion paper. Mar. Environ. Res. 96, 81–91. http://dx.doi.org/ 10.1016/j.marenvres.2013.10.008. Bradford, M.M., 1976. A rapid and sensitive method for the quantitation of microgram quantities of protein utilizing the principle of protein-dye binding. Anal. Biochem. 7, 248–254. http://dx.doi.org/10.1016/0003-2697(76)90527-3. Broeg, K., Lehtonen, K.K., 2006. Indices for the assessment of environmental pollution of the Baltic Sea coasts: integrated assessment of a multi-biomarker approach. Mar.

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