Science of the Total Environment 543 (2016) 147–154

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Toxicological and chemical assessment of arsenic-contaminated groundwater after electrochemical and advanced oxidation treatments Sandra Radić a,⁎, Helena Crnojević b, Valerija Vujčić a, Goran Gajski c, Marko Gerić c, Želimira Cvetković d, Cvjetko Petra a, Vera Garaj-Vrhovac c, Višnja Oreščanin e a

University of Zagreb, Faculty of Science, Department of Biology, HR-10 000, Croatia Croatian Waters, Main Water Management Laboratory, HR-10 000 Zagreb, Croatia Institute for Medical Research and Occupational Health, Mutagenesis Unit, HR-10 000 Zagreb, Croatia d Institute of Public Health “Dr. Andrija Štampar”, Department of Environmental Protection and Health Ecology, HR-10000 Zagreb, Croatia e ORESCANIN Ltd., HR-10 000 Zagreb, Croatia b c

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• Toxicity of treated and untreated arsenic-contaminated groundwater was assessed. • Chemical analysis and bioassays on human cells, plants and Daphnia were conducted. • Genotoxic and toxic compounds induced oxidative stress. • Toxicity and genotoxicity were removed following the treatment. • Chronic toxicity test more suitable in evaluation of arsenic-polluted groundwater

a r t i c l e

i n f o

Article history: Received 21 May 2015 Received in revised form 30 October 2015 Accepted 30 October 2015 Available online xxxx Editor: D. Barcelo Keywords: Arsenic Toxicity Human lymphocytes

a b s t r a c t Owing to its proven toxicity and mutagenicity, arsenic is regarded a principal pollutant in water used for drinking. The objective of this study was the toxicological and chemical evaluation of groundwater samples obtained from arsenic enriched drinking water wells before and after electrochemical and ozone–UV–H2O2-based advanced oxidation processes (EAOP). For this purpose, acute toxicity test with Daphnia magna and chronic toxicity test with Lemna minor L. were employed as well as in vitro bioassays using human peripheral blood lymphocytes (HPBLs). Several oxidative stress parameters were estimated in L. minor. Physicochemical analysis showed that EAOP treatment was highly efficient in arsenic but also in ammonia and organic compound removal from contaminated groundwater. Untreated groundwater caused only slight toxicity to HPBLs and D. magna in acute experiments. However, 7-day exposure of L. minor to raw groundwater elicited genotoxicity, a significant growth inhibition and oxidative stress injury. The observed genotoxicity and toxicity of raw groundwater samples was almost

⁎ Corresponding author at: University of Zagreb, Faculty of Science, Department of Biology, Rooseveltov trg 6, HR-10 000 Zagreb, Croatia. E-mail address: [email protected] (S. Radić).

http://dx.doi.org/10.1016/j.scitotenv.2015.10.158 0048-9697/© 2015 Elsevier B.V. All rights reserved.

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Lemna minor Daphnia magna Advanced treatment

completely eliminated by EAOP treatment. Generally, the results obtained with L. minor were in agreement with those obtained in the chemical analysis suggesting the sensitivity of the model organism in monitoring of arseniccontaminated groundwater. In parallel to chemical analysis, the implementation of chronic toxicity bioassays in a battery is recommended in the assessment of the toxic and genotoxic potential of such complex mixtures. © 2015 Elsevier B.V. All rights reserved.

1. Introduction

33″ E) and Vrbanja (47°14′5″ N, 6°31′17″ E) – situated in the Vukovar-Srijem County, Eastern Croatia were used as sampling sites. As much as 300 L of groundwater was collected per well and stored at +4 °C. Water was pumped into acid cleaned polyethylene containers using a handheld electric deep water sampler, model ZYC-2A (EJER TECH, Zhejiang, China). Homogenization of water samples was performed for 10 min before the treatments and analysis using mixer model SJBQ-2.2 (Siehe Industry, Shanghai, China).

Inorganic arsenic species are able to cause cancer in humans though it is harmful to other biota as well (Ng et al., 2003). In many nations, more than half of the withdrawn groundwater is used for domestic water supplies providing 25 to 40% of the world's drinking water (WWAP, 2006). The main environmental exposure to elevated concentrations of arsenic for humans is through contaminated groundwater, imposing a health risk for millions of people worldwide (Ng et al., 2003). Uncontaminated freshwaters seldom contain more than 10 μg L−1 of arsenic; this concentration has been determined as a threshold for arsenic in drinking water by the World Health Organization (WHO) (Sharma and Sohn, 2009). Although in most European countries this standard is rarely exceeded, elevated concentrations of naturally-occurring arsenic in groundwater are found in the Pannonian Basin (Eastern Europe). This incident may cause serious health problems upon prolonged exposure for more than half a million people in several countries of Eastern Europe (Van Halem et al., 2009). Approximately 200,000 inhabitants of Eastern Croatia are exposed to arsenic at levels ranging from 10 to 610 μg L−1 since the groundwater from water wells is used as a principal drinking-water source (HabudaStanić et al., 2007). Various purification methods have been developed for the treatment of arsenic contaminated water (Choong et al., 2007). Besides arsenic, the groundwater in Eastern Croatia is often loaded with high concentrations of iron, manganese, ammonia and organic substances (Ujević et al., 2010). In the present study, arsenic-contaminated groundwater samples collected from four drinking water-wells in Vinkovci County (Eastern Croatia) were subjected to simultaneous electrocoagulation (using iron- and aluminum-electrodes) and ozonation and subsequent advanced oxidation processes (AOP) comprising simultaneous ozone and UV light and/or hydrogen peroxide treatment (Oreščanin et al., 2013; Oreščanin et al., 2014). The aim of this research was to evaluate the acute and chronic toxicological effects of groundwater samples obtained from drinking water wells prior to and after the treatment. For this purpose, two standardized toxicity tests (ISO 6341, 2012; ISO 20079, 2005) with freshwater organisms L. minor and D. magna were applied. Since the contaminated groundwater is still commonly used for human consumption, human cells were also employed as target organisms. Bioassays (cell viability and comet assays) with peripheral blood lymphocytes (HPBLs) used as sensitive in vitro models have previously proven their suitability in the evaluation of cytotoxic and genotoxic potential of arsenic and carcinogenic metals (Gajski et al., 2015). The use of a battery of bioassays with model systems covering different trophic levels is recommended in water quality testing as toxic substances have different modes of action and target receptors and thus do not produce the same effects in divergent test organisms (Hernández Leal et al., 2012; Kungolos et al., 2015). Since L. minor showed high sensitivity to raw groundwater in our preliminary experiments, additional analysis (comet assay and indicators of oxidative stress) was performed with this model organism. Oxidative stress parameters were analyzed as the number of contaminants, including arsenic and metals, have been found to act as potent elicitors of reactive oxygen species (ROS) formation (Fodor, 2004; Radić et al., 2013). 2. Material and methods 2.1. Collection and handling of water samples Drinking wells from four locations – Andrijaševci (46°1′27″ N, 6°48′ 35″ E), Antin (44°43′51″ N, 6°59′29″ E), Komletinci (47°5′50″ N, 6°25′

2.2. Purification experiments 2.2.1. Removal of arsenic and metals All purification experiments were conducted at room temperature (22 °C). The method was described in detail in Oreščanin et al. (2014). In short, a certain volume of groundwater from each of the four drinking water wells (each water sample was taken in triplicate) was treated with ozone (OzoneMax 1668, Ozonemax Water Technologies, Kochi, Kerala, India) for 10 min to convert arsenite to arsenate as the latter shows the highest adsorption capacity for both iron- and aluminumelectrodes. Electrocoagulation was first performed with an iron- and then with an aluminum-electrode set, in each case with parallel arrangement — 12 cathodes and 12 anodes (dimension of individual electrode 200 × 500 × 2 mm, separation among electrodes 5 mm, reaction time 5 min). The following steps included mixing the solution with ozone (10 min), passing through an electromagnet and collection of the suspension in sedimentation tank where solids were separated from liquid (60 min). The treatment unit was a patented product (WO2013144664A9) of Advanced energy Ltd., Zagreb, Croatia. The unit and treatment method were also described in detail in our previous work (Oreščanin et al., 2013). 2.2.2. Ammonia and organic matter removal Further treatment of partially treated groundwater samples included UV light (60 W) and ozone. In the case of AS and KO groundwater samples, 0.15 mL L−1 of 30% hydrogen peroxide was also added to the solution and treated for 30 min while the other two samples (AT and V) were treated for 30 min with UV and ozone only. 2.3. Physicochemical and toxicological analysis Water samples for physicochemical and toxicological evaluation were collected from the production wells (raw groundwater samples: AS, AT, KO, V) and from the treatment plant after the completion of the purification process (AST, ATT, KOT, VT). 2.3.1. Physicochemical analysis Assessment of water quality is done in accordance with international standards (ISO) by authorized laboratories. pH and conductivity (mS cm−1) of raw (untreated) groundwater samples were determined on site whereas both parameters of all of the treated groundwater samples were measured in a laboratory. Chemical analyses included color, turbidity, suspended solids (SS, mg L−1), chemical oxygen demand (COD, mg of O2 per liter), total organic carbon (TOC), nitrate (mg L−1), nitrite (mg L−1), total ammonia −1 ), cloride (Cl−) and fluoride (mg L−1), soluble phosphate (PO3− 4 , mg L − (F ) contents (EN 1484, 1997; ISO 7888, 1985; ISO 6060, 1989; ISO 10523, 1994; ISO 11923, 1997; ISO 14911, 1998; ISO 7027, 1999; ISO 10304, 2007; ISO 7887, 2011; ISO/TR 11905, 1997). Mercury was determined by atomic fluorescent spectrometry (QuickTrace M-8000

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Mercury Analyzer, Teledyne Cetac Technologies, Omaha, NE) according to EN ISO17852 (2008) and ISO 17852 (2006). Cadmium, lead, chromium, iron, manganese, cobalt, nickel, copper, zinc, vanadium and arsenic were detected using inductively coupled plasma mass spectrometry (Perkin Elmer ELAN 9000, Waltham, MA) according to EN ISO 172942 (2004) and ISO 17294-2 (2003). Hydrogen peroxide was determined by 110011 Peroxide Test (MQuant™, Merck Millipore, Darmstadt, Germany). 2.3.2. Cytotoxicity (cell viability) and genotoxicity (comet assay) in human lymphocytes The effects of the untreated/treated samples were assessed in HPBLs obtained from a young healthy male, non-smoking donor who had not been exposed to diagnostic or therapeutic radiation or known genotoxic chemicals that might have interfered with the results of the testing for a 12 months before blood sampling. The study was part of a project approved by the Institutional Ethics Committee and observed the ethical principles of the Declaration of Helsinki. For toxicological testing, each groundwater sample was added to peripheral blood to make the final concentration. The whole blood was used for the comet assay, while cytotoxicity was assessed on isolated HPBLs. The groundwater samples were added to the whole blood samples during 4 and 24 h period. In each experiment, a non-treated control was included which consisted of an equal amount of sterile distilled water opposite to arseniccontaminated groundwater which was used for the treated samples. Cytotoxicity was determined by differential staining with acridine orange and ethidium bromide and by fluorescence microscopy (Duke and Cohen, 1992). A total of 100 cells per repetition were examined with an epifluorescent microscope (Olympus BX51, Tokyo, Japan) and the results were expressed as percentage of viable cells. The alkaline comet assay was done according to the standard protocol as described by Singh et al. (1988). Afterwards, the slides were stained with ethidium bromide and analyzed using an epifluorescence microscope (Zeiss, Göttingen, Germany) with the image analysis system (Comet Assay II; Perceptive Instruments Ltd., Haverhill, Suffolk, UK). One hundred randomly captured comets were analyzed from each slide and the results were expressed as % of tail DNA. 2.3.3. Daphnia magna immobilization test D. magna Straus was provided by the Laboratory for Environmental Toxicology, Ghent University, Belgium. Tests were carried out with neonates hatched from dormant eggs (Daphtoxkit F™ Magna, 1996; ISO 6341, 2012). The ephippia were incubated for 72 h, at 20 ± 2 °C under continuous illumination of 80 μEm−2 s−1. Dilution series (1–100%) of groundwater samples before and after treatment were made. The dilution medium for either dilution series of groundwater samples or control was prepared according to the ISO 6341 (2012). Five young daphnids (b 24 h) were placed into test tubes containing 10 mL different dilutions (each with four replicates) of the groundwater samples. The test was run at temperatures of 20 ± 2 °C in the dark. Following 24 h or 48 h, the immobilization was determined. EC50 values were calculated by use of regression line and the values were then expressed as toxic units (TU = 1/EC50) according to which TU b 0.4 = no acute toxicity, 0.4 b TU b 1 = slight acute toxicity, 1 ≤ TU b 10 = acute toxicity, 10 ≤ TU b 100 = high acute toxicity and TU ≥ 100 = very high acute toxicity (Persoone et al., 2003). 2.3.4. Lemna test and toxicity parameters in duckweed The axenic stock cultures of L. minor were maintained on Pirson–Seidel medium (Pirson and Seidel, 1950) in a growth chamber at 24 ± 2 °C, 16 h photoperiod and photon flux density of 90 μEm−2 s−1 (TLD 36 W/ 54-765; Philips, Poland). Pre-cultivation (two-week adaptation period) and cultivation (a week period) of plants for Lemna test was achieved on Steinberg nutrient media defined by ISO 20079 test protocol (ISO 20079, 2005) and over that time plants were cultivated in a growth chamber at 24 ± 2 °C, under continuous light and photon flux density of 90 μEm− 2 s− 1. Uniform, healthy colonies with 2–3 fronds (from

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stock cultures) were cultivated in test beakers (100 cm3) filled with 60 cm3 of Steinberg nutrient media containing either raw (untreated) or purified (treated) groundwater samples from four drinking wells (AS, AT, KO, V, AST, ATT, KOT, VT). Control plants were cultivated on Steinberg nutrient media prepared with distilled water. Before experiment, groundwater samples were passed through nitrocellulose membranes (Whatman, 0.45 μm). All analyses were performed with 9 replicates. The experiment lasted 7 days. Frond number (FN) and fresh weight (FW) were used as biomass parameters (x) for estimation of growth rate (RGR). The parameters were noted at the beginning (t1) and at the end of the Lemna test (t2). RGR was estimated using the equation: (ln xt2 − ln xt1)/t2 − t1. The chlorophyll a level was determined in centrifuged 80% (v/v) acetone extracts and estimated according to Lichtenthaler (1987). The formation of malondialdehyde (MDA) and thiobarbituric acid adduct was used as a measure of lipid peroxidation (Heath and Packer, 1968). The carbonyl groups (C_O) formed in the reaction with 2,4-dinitrophenylhydrazine (Sigma-Aldrich) were used as indicators of protein oxidation (Levine et al., 1990). For analysis of antioxidant enzymes, plant tissue was homogenized in potassium phosphate buffer (50 mM) with pH 7 containing polyvinylpolypyrrolidone (PVPP) and EDTA (1 mM) using prechilled mortars and centrifuged for 30 min at 4 °C and 25,000 g. Superoxide dismutase (SOD) and catalase (CAT) activities were determined according to Beauchamp and Fridovich (1971) and Aebi (1984), respectively. The activity of both enzymes was presented as unit per mg protein. DNA damage to duckweed nuclei was detected by comet assay (Gichner et al., 2004). Minor changes were introduced to the protocol: denaturation 10 min, electrophoresis 20 min, 1 V cm−1). For each groundwater sample, three replicates with 100 nuclei per each replicate were analyzed by a computerized image-analysis system (Komet version 5, Kinetic Imaging Ltd., Liverpool, UK) linked to the fluorescence microscope (excitation filter BP 520/09 nm, barrier filter 610 nm). The comet parameter for estimation of DNA damage was as % of tail DNA. ROS levels in duckweed cells were determined using a superoxide anion specific probe, dihydroethidium (DHE) (Sandalio et al., 2008). Duckweed fronds were incubated with DHE in Tris–HCl buffer (10 mM, pH 7.4) and subsequently washed to remove the residual dye. Superoxide oxidizes DHE to oxyethidium, a stable red fluorescent product. The slides were examined under epifluorescent microscope Olympus BX-51 connected to the camera (Olympus DP70, Tokyo, Japan) at excitation 450–490 nm and emission 520 nm or more. Fluorescence intensity was measured and images were analyzed using software Lucida 6.0 (Wirral, UK). Three slides were evaluated per water sample. A total number of at least 100 cells were counted in 25 fields of each slide. Identical conditions were kept in all of the experimental groups. 2.3.5. Statistical analyses Statistical analysis was done by STATISTICA 12.0 (StatSoft, Inc., USA). Normality of the data was tested by Shapiro–Wilk's W test. Homogeneity of variance for each dependent variable was tested by Levene's Test. The possible difference among the nine groups of samples: negative control (C), each raw water well sample before the treatment (AS, AT, KO, V) and the same samples after the passing water treatment plant (AST, ATT, KOT, VT) was assessed by one-way ANOVA followed by Newman–Keuls post hoc comparison test. Since the results of the comet assay obtained on the human blood lymphocytes (for both exposure periods) deviated from the normal distribution, these data were logtransformed prior to the ANOVA and Newman–Keuls test. All other data showed normal distribution and required no further treatment. In all of the statistical tests, the significance level was set to (P b 0.05). 3. Results and discussion In this study, toxicological and chemical analysis was used to monitor the effects of micropollutants present in groundwater prior to and

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following EAOP treatment. The physicochemical parameters of four groundwater samples from the drinking water wells prior/after the treatment are presented in Table 1 together with national limit values which, for most of the parameters, agree with guideline limits set by the WHO (WHO, 2011). The results reveal that most parameters, except arsenic in all water samples and a few parameters in certain water samples, do not exceed the permissible limit prescribed by national maximum allowed values (MAV). Color, turbidity and COD exceeded MAV in the groundwater sample AS which together with TOC values of the sample indicated a certain amount of dissolved oxidizable organic substances including non-biodegradable matter. Color and ammonia nitrogen increased above MAV in V and KO water samples, respectively. Of all the metals measured, only manganese exceeded the national permissible limit in several samples (AT, V and AS), though the guideline value for manganese in drinking water set by the WHO is several time higher (400 μg L−1) compared to the limit set for Croatian surface and drinking water (50 μg L−1). Arsenic exceeded the upper permissible limit (10 μg L−1) around 18 and 14 times in AS and KO water samples respectively, whereas the increase of the metalloid in the two other water samples was nearer to MAV. Higher arsenic concentrations in the two water samples (AS and KO) are comparable to average values (greater than 100 μg L−1) of the metalloid recorded in well water from neighboring arsenic-affected countries (Varsányi and Kovács, 2006; Oreščanin et al., 2013). However, in a previous investigation by Ćavar et al. (2005), more than 600 μg L−1 of arsenic was detected in groundwater from the Andrijaševci water well. Following electrocoagulation with ozonation and subsequent ozonolysis and simultaneous oxidation with ROS formed by UV light and ozone, the majority of the physicochemical parameters decreased and were well below the regulated values. Increased ammonia and COD from the two water-wells (AS and KO) were removed by additional oxidation with hydrogen peroxide. Arsenic triggers carcinogenesis by several mechanisms — induction of DNA damage, interference with DNA repair, cell cycle and differentiation, induction of DNA hypomethylation (and thus genomic instability), modulation of signal transduction pathways (which subsequently affects cell growth, apoptosis and cell adhesion) and overproduction

of ROS (Shi et al., 2004; Sun et al., 2014). The arsenic-induced cytotoxicity and genotoxicity is essentially explained by the said mechanisms (Sun et al., 2014). Chronic exposure (several years) of humans to relatively high arsenic concentrations (between 100 and 500 mg L−1) in drinking water induced cytotoxicity effects, ROS-induced damage to vital biomolecules and DNA damage in HPBLs (Basu et al., 2005). In the current study, the genotoxic activity of untreated groundwater samples to HPBLs was not detected by comet assay (Fig. 1). Regardless of the exposure period (4 and 24 h), DNA damage in HPBLs exposed to either untreated or treated groundwater samples was similar to values noted in negative control. Unlike genotoxic, the cytotoxic activities obtained after 4 and 24 h show that the percentage of the viable cells decreases with increasing exposure time (Fig. 1). All untreated groundwater samples, irrespective of their arsenic content, caused a small but significant decline in cell viability in comparison to controls over both exposure periods. The results are comparable to prior studies demonstrating an absence of or mild genotoxic and cytotoxic effects upon short-term (up to 24 h) exposure of cultured human cell lines to arsenic (Hartmann and Speit, 1994; Chattopadhyay et al., 2002). The mild cytotoxic effect observed in acute experiments with HPBLs cannot be ascribed to any specific toxic action but rather to the joint effects of micropollutants present in untreated groundwater, including arsenic and metals. Regardless of exposure period, the cell survival of HPBLs exposed to treated groundwater was similar to control. D. magna acute immobilization test proved sensitive enough and thus relevant for evaluation of groundwater toxicity (Dewhurst et al., 2002). In our study however, similarly as with short-term exposure of HPBLs, the results of acute D. magna test showed the necessity for longer-term tests in the assessment of arsenic-enriched groundwater toxicity. To be more precise, after the exposure of daphnids to untreated groundwater samples only a slight toxicity was detected in our study — the TU values were 0.59 and 0.73 for untreated V and KO, respectively, after 24 h while the TU values for untreated water samples ranged from 0.63 to 0.86 after 48 h of exposure (Table 2). The highest toxicity to D. magna was caused by untreated KO — 40% inhibition of immobilization at 48 h. In prior reports, a very high acute (48 h) or chronic

Table 1 Physicochemical parameters in the contaminated groundwaters before and after the electrochemical treatment and maximum allowed values (MAV) in accordance with national regulations. Groundwater samples

AS

Measured parameter

Before EAOP treatment

AT

Color (PtCo) Turbidity (NTU) pH SS (mg L−1) EC (mS cm−1) CODCr (mg L−1) TOC H2O2 (mg L−1) −1 ) NO− 3 (mg L −1 NO− ) 2 (mg L −1 + NH4 (mg L ) −1 3− PO4 (mg L ) F− (mg L−1) Cl− (mg L−1) Na+ (mg L−1) As (μg L−1) Hg (μg L−1) Cd (μg L−1) Pb (μg L−1) Cr (μg L−1) Fe (μg L−1) Mn (μg L−1) Co (μg L−1) Ni (μg L−1) Cu (μg L−1) Zn (μg L−1) V (μg L−1)

76 12 7.81 5 0.61 8.48 2.99 0 0.039 0.148 0.199 0.103 0.105 11.02 140.79 179 0.006 0.048 0.135 6.23 82.1 54.9 0.194 1.07 1.39 2.64 0.402

12 0 7.33 0 0.66 2.94 0.839 0 0.005 0.002 0.250 0.090 0.397 3.26 72.14 15.15 0.002 0.03 0.086 6.04 79.6 140.0 0.175 0.279 2.36 2.59 0.331

KO

V

AST

ATT

KOT

VT

0 0 7.92 0 0.04 b2 0.114 0.5 0.008 0.002 0.08 0.0033 0.001 0.311 5.11 b0.03 b0.001 0.014 0.082 1.27 b0.03 0.067 b0.008 0.048 0.731 1.1 b0.03

0 0 7.74 0 0.07 b2 0.106 0 0.002 b0.001 0.034 0.0027 0.003 0.192 9.08 b0.03 0.004 0.014 0.092 0.19 b0.03 0.166 b0.008 0.099 1.27 1.13 b0.03

After EAOP treatment 10 0 8.1 0 0.49 2.94 1.03 0 0.021 0.009 0.673 0.199 0.03 1.88 62.59 138 0.001 0.021 0.17 5.27 45.0 10.1 0.057 0.079 1.14 2.04 0.327

31 0 7.56 0 0.78 3.91 1.12 0 0.034 0.002 0.298 0.147 0.541 2.81 134.42 12.50 0.009 0.044 0.126 22.30 3.1 154.5 0.113 0.354 3.41 1.37 0.36

0 0 7.21 0 0.06 b2 0.162 0.5 0.011 0.004 0.011 0.0024 0.002 0.303 9.95 b0.03 b0.001 0.019 0.093 0.802 0.773 0.099 b0.008 0.112 0.88 1.3 b0.03

0 0 7.43 0 0.06 b2 0.085 0 b0.001 b0.001 0.005 0.0024 0.002 0.108 3.67 b0.03 b0.001 0.011 0.08 b0.03 b0.03 0.177 b0.008 0.097 1.3 1.13 b0.03

National MAV 20 4 6.5–9.5 10 2.5 5 ND ND 50 0.1 0.5 0.3 1.5 250 150 10 1 5 10 50 300 50 ND 20 2000 3000 5

SS — suspended solids, EC — electrical conductivity, COD — chemical oxygen demand, TOC — total organic carbon. ND - not determined. Numbers are means of two replicates.

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Fig. 1. A) Cell viability (% of control), and B) % of DNA in tail (comet assay parameter) of human peripheral blood lymphocytes (HPBLs) following 4 and 24 h exposure periods to untreated (original) and treated groundwater samples. C — control; AS — untreated groundwater sample from Andrijaševci; AT — untreated groundwater sample from Antin; KO — untreated groundwater sample from Komletinci; V — untreated groundwater sample from Vrbanja; AST — treated groundwater sample from Andrijaševci; ATT — treated groundwater sample from Antin; KOT — treated groundwater sample from Komletinci; VT — treated groundwater sample from Vrbanja. Standard deviations are presented by error bars. Different letters indicate significantly different values at P b 0.05.

(21 days) toxicity was determined for Daphnia sp. upon arsenic exposure, but in both studies extremely high arsenic concentrations were used (Hoang et al., 2007; Shaw et al., 2007). Here, the concentrations of arsenic as well as of other micropollutants in raw groundwater samples were probably too low (well below those associated with mortality) to exert their toxic effect on D. magna over relatively short-term exposure. Regarding the efficiency of EAOP treatment, the toxicity of KO water sample was reduced partially and of V water sample completely. Other treated groundwater samples showed no acute toxicity. Overall, the results obtained on duckweed suggest that the test system is highly sensitive for the monitoring of arsenic-contaminated waters. Certain indicators – levels of ROS, parameters of oxidative damage to lipids and proteins, parameters of comet assay – proved to be more responsive to changes in arsenic content in the groundwater samples Table 2 The effects of untreated and treated groundwater samples from 4 water-wells to D.magna following 24 and 48 h exposure. Daphnia toxicity TU Untreated water AS AT KO V Treated water AST ATT KOT VT

24 h

48 h

b0.4 b0.4 0.73 0.59

0.83 0.75 0.86 0.63

b0.4 b0.4 0.51 b0.4

b0.4 b0.4 0.73 b0.4

and thus more suitable for the evaluation of arsenic-induced toxicity. The relative growth rate, expressed either by frond number or fresh weight basis, was significantly inhibited by all groundwater samples though the higher inhibition of RGR corresponded with higher arsenic content detected in water samples AS and KO (Fig. 2a). Similar results were noted with chlorophyll a; the content of the pigment declined to 58% of control in AS and KO groundwater samples and to 67% of control in AT and V groundwater samples (Fig. 2b). A similar pattern of changes to water samples was observed with chlorophyll b and carotenoid contents (data not shown). A decrease in photosynthetic pigment content and then often subsequent growth retardation is one of the early responses to many stress agents such as metals and metalloids. The harmful effects of arsenic and metals on photosynthetic pigments is usually attributed to the inactivation of enzymes involved in chlorophyll synthesis or to peroxidation of chloroplast membranes as a consequence of ROS overproduction (Fodor, 2004). Here, taking into account the results of chemical analysis, there is no clear connection between arsenic and the observed negative effects of raw groundwater samples on growth and photosynthetic pigments; thus the observed toxicity could be attributed to cumulative toxic action of the groundwater constituents including arsenic (Leão et al., 2014) rather than to specific toxic effects of arsenic alone. The results of oxidative stress parameters imply oxidative damage to membrane lipids and proteins; ROS, MDA and C_O levels markedly increased in response to all groundwater samples, though especially to those with higher arsenic content (Fig. 2d, e, f). There is significant evidence that arsenic exerts its toxicity by reacting with thiol groups and competing with phosphate (as it acts as a phosphate analogue) in cellular processes, though the major toxic action of arsenic is considered to be redox stress and the related increased generation of ROS (Meharg

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Fig. 2. A) Relative growth rate, B) chlorophyll a (mg g−1 FW), C) % of DNA in tail (comet assay parameter), D) MDA — indicator of lipid peroxidation (nmol g−1 FW), E) ROS (% of control), F) C_O — indicator of protein oxidative damage (nmol mg−1 protein), G) SOD activity (U mg−1 protein), and H) CAT activity (U mg−1 protein) of duckweed following 7-day exposure period to untreated (original) and treated groundwater samples. C — control; AS — untreated groundwater sample from Andrijaševci; AT — untreated groundwater sample from Antin; KO — untreated groundwater sample from Komletinci; V — untreated groundwater sample from Vrbanja; AST — treated groundwater sample from Andrijaševci; ATT — treated groundwater sample from Antin; KOT — treated groundwater sample from Komletinci; VT — treated groundwater sample from Vrbanja. Standard deviations are presented by error bars. Different letters indicate significantly different values at P b 0.05.

and Hartley-Whitaker, 2002; Hasanuzzaman et al., 2015). The latter mechanism – direct or indirect ROS overproduction via disturbance of redox status or interference with the plant's antioxidant defense system – is partly responsible for metal toxicity as well (Fodor, 2004). It has been well-documented that both arsenic and other stressors can upregulate antioxidative enzyme activities (Fodor, 2004, Hasanuzzaman et al., 2015) which was confirmed in our study though only partially (Fig. 2g). The activity of SOD, the first enzyme of the antioxidative defense system

which dismutates superoxide to hydrogen peroxide, was markedly enhanced in plants exposed to groundwater samples, and again the increase was far greater in response to groundwater samples with higher arsenic content (Fig. 2g). However, the activity of CAT (Fig. 2h) as well as ascorbate peroxidase (data not shown), hydrogen peroxide degrading enzymes, was significantly inhibited. The activity of CAT dropped to 25 and 35% of control in AS and KO groundwater samples, respectively and to 43 and 49% of control in AT and V groundwater

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samples, respectively. A similar result, the enhancement of SOD and inhibition of H2O2-destroying enzymes was recorded in Lemna gibba exposed to arsenic (Leão et al., 2014). Apart from the destructive effects on lipids and proteins, all raw groundwater samples, especially those with higher arsenic concentration (AS and KO), induced genotoxicity to duckweed nuclei as evident from the results of the comet assay (Fig. 2c). Groundwater samples AT and V increased DNA damage (% of DNA in tail) between 23 and 36%, whereas AS and KO samples caused a much higher degree of DNA damage (between 85 and 95%) with respect to control. Genome damage caused by arsenic and metals may be induced by elevated generation of ROS whereas arsenic was also reported to interfere with DNA repair enzymes (Patra et al., 2004; Lin et al., 2008). Regarding the results of oxidative stress parameters and of comet assay, the observed biological effects corresponded with the arsenic concentrations in raw groundwater samples though other pollutants such as ammonia, COD or manganese in certain samples might have contributed to such effects as well. In contrast to genotoxicity, EAOP treatment did not completely reduce the toxicity of all groundwater samples to L. minor. According to the results, some residual toxicity remained in AS and KO groundwater samples since the ROS and CAT levels of the treated samples maintained significantly higher values than control (Fig. 2e, g, h). The reason for the observed toxicity may lie in the additional oxidation of the samples with hydrogen peroxide (Table 1) which was performed to remove COD and ammonia from the AS and KO samples, respectively. However, damage to vital biomolecules (DNA, proteins and lipids) of duckweed was completely reduced by EAOP treatment. 4. Conclusions The obtained results suggest the utility of chronic (long-term) rather than acute (short-term) toxicity bioassays as more suitable in the toxicity assessment of the arsenic-contaminated groundwater. Experiments with L. minor revealed a significant toxicity and genotoxicity in raw groundwater which was almost completely reduced after EAOP treatment. The results provided by chemical analysis were mostly consistent with the results obtained with L. minor — certain indicators (oxidative stress and genotoxicity parameters) proved to be relatively sensitive to arsenic concentrations in groundwater and thus potentially useful in the evaluation of arsenic-induced toxic effects. Conflict of interest The authors declare that there are no conflicts of interest. Acknowledgments This study was funded by Hrvatske Vode (Croatian Waters) as part of the project “The evaluation of quality of surface- and groundwater using biotests”. References Aebi, H., 1984. Catalase in vitro. Methods Enzymol. 105, 121–126. Basu, A., Som, A., Ghoshal, S., Mondal, L., Chaubey, R.C., Bhilwade, H.N., et al., 2005. Assessment of DNA damage in peripheral blood lymphocytes of individuals susceptible to arsenic induced toxicity in West Bengal, India. Toxicol. Lett. 159 (1), 100–112. Beauchamp, C., Fridovich, I., 1971. Superoxide dismutase: improved assay and an assay applicable to PAGE. Anal. Biochem. 44, 276–287. Chattopadhyay, S., Bhaumik, S., Chaudhury, A.N., Gupta, S.D., 2002. Arsenic induced changes in growth development and apoptosis in neonatal and adult brain cells in vivo and in tissue culture. Toxicol. Lett. 128 (1–3), 73–84. Choong, T.S.Y., Chuah, T.G., Robiah, Y., Gregory Koay, F.L., Azni, I., 2007. Arsenic toxicity, health hazards and removal techniques from water: an overview. Desalination 217, 139–166. Daphtoxkit F™ magna, 1996. Crustacean toxicity screening test for freshwater. Standard Operational Procedure (Creasel, Deinze, Belgium).

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