Effects of benthic organism Tubifex tubifex on hexachlorocyclohexane isomers transfer and distribution into freshwater sediment

Ecotoxicology and Environmental Safety 126 (2016) 163–169

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Effects of benthic organism Tubifex tubifex on hexachlorocyclohexane isomers transfer and distribution into freshwater sediment Shanshan Di a,b, Cheng Cheng b, Li Chen a,b, Zhiqiang Zhou a,b, Jinling Diao b,n a b

Beijing Advanced Innovation Center for Food Nutrition and Human Health, China Agricultural University, Yuanmingyuan west road 2, Beijing 100193, China Department of Applied Chemistry, China Agricultural University, Yuanmingyuan West Road 2, Beijing, 100193, China

art ic l e i nf o

a b s t r a c t

Article history: Received 20 September 2015 Received in revised form 7 December 2015 Accepted 15 December 2015

In this study, bioaccumulation and elimination of HCHs in tubifex, and the distribution of HCHs in overlying water and sediment, were studied during a 10-d experiment. A sensitive method was developed for the determination of HCHs in samples based on gas chromatograph (GC) equipped with a nickel-63 electron capture detector (μECD). The limit of detection (LOD) was 0.35 mg/kg for α-HCH and 0.82 mg/kg for β-HCH. Tubifex accumulated HCHs rapidly, and the curves were approximately M-type. The highest level was reached on the 7th day, with 0.34 mg/kgwwt for α-HCH and 0.87 mg/kgwwt for βHCH in worms. The AFs of β-HCH in tubifex were higher than those of α-HCH. Moreover, the existence of tubifex significantly reduced β-HCH fluxes from the overlying water to sediment by uptake or degradation and decreased the concentrations of β-HCH in the sediment, but it had little influence on αHCH fluxes. Moreover, enantioselectivity of α-HCH enantiomers was not observed in tubifex, whether in the bioaccumulation or elimination experiments. At the end of the elimination experiment, approximately 80% and 70% of α-HCH and β-HCH were eliminated, and the depuration half-lives were 4.43 and 5.39 days, respectively. & 2015 Elsevier Inc. All rights reserved.

Keywords: Tubifex tubifex HCHs Bioaccumulation Elimination Sediment

1. Introduction Organochlorine pesticides (OCPs) have attracted wide attention for decades due to their large production and usage, their nature of persistence, bioaccumulation and harmful effects in the environment (Ali and Jain, 1998; Hussain et al., 2015; Jones and De Voogt, 1999; Nakata et al., 2002; Zhou et al., 2013). Although most of them have been officially banned for many years, OCPs residues are still detectable in various environmental media, even in remote high latitude or high altitude regions, such as polar regions, Tibetan Plateau and high mountains (Zhang et al., 2013; Zheng et al., 2009). Most organochlorine pesticides are categorized as persistent organic pollutants (POPs), and hexachlorocyclohexanes (HCHs) were recognized as POPs by the Stockholm Convention in 2009 (UNEP). HCHs were used as worldwide insecticides during the 1950s to 1980s (Ali and Ali, 2004; Niu et al., 2013). Two types of HCHs products were used around the world: technical HCHs (mainly containing α-HCH, β-HCH, and γ-HCH) and lindane (γHCH 499%) (Zhang et al., 2011). These HCH isomers have different physicochemical properties, which cause different partitioning in n

Corresponding author. E-mail address: [email protected] (J. Diao).

http://dx.doi.org/10.1016/j.ecoenv.2015.12.022 0147-6513/& 2015 Elsevier Inc. All rights reserved.

the environment (Wu et al., 2013). α–HCH has a high vapor pressure and is prone to evaporation to the atmosphere. In contrast, β-HCH, with all chlorine atoms in the equatorial conformation, is in the most energetically favorable configuration, therefore it is the most stable molecule among the HCH isomers (Chessells et al., 1988). If there were no fresh inputs of technical HCH, β-HCH would be the predominant isomer in most sediments (Wu et al., 2013). OCPs have a strong affinity for suspended particulates and sediments, on account of their low-water solubility and high n-octanol/water partitioning coefficient values (Kow). Thus, sediments can serve as reservoirs, or “sinks” for OCPs (Yang et al., 2011). Once disturbed, the sediments can be resuspended and the contaminants may re-enter the aquatic environment, resulting in a second contamination (Yang et al., 2011; Zeng and Venkatesan, 1999). Benthic organisms, living at the sediment-water interface, transfer the sediment-associated contaminants to mix back into the aqueous phase by bioturbation, and/or accumulate sedimentary pollutants and subsequently transfer them into higher trophic levels through the food web (Egeler. et al., 2001; Liu et al., 2015). Sediment-associated POPs are known to exhibit narcotic effects in benthic organisms, and they also have been implicated in the development of tumors, malformation, loss of fertility, or immune deficiency in many organisms (Liu et al., 2009; Lu et al., 2013). Cosmopolitan aquatic oligochaete, tubifex worms have intimate

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contact with the aqueous and solid phase, burrowing the anterior part into the sediment and undulating the posterior part in the overlying water to allow cutaneous respiration (Liu et al., 2015). Tubifex can be found in both unpolluted and highly polluted waters and is often the last to disappear in a contaminated site (Reynoldson et al., 1996). Previous studies have shown that the autotomy and regeneration of their caudal part may participate in efficient detoxification of tubifex, which make them survive in highly contaminated media (Lagauzere et al., 2009; Meller et al., 1998). Moreover, T. tubifex plays an important role in self-purification of water bodies, and has been designated as a representative freshwater test organism to assess sediment toxicity and accumulation of organic compounds (Fangtong et al., 2011; Lagauzere et al., 2009; Liu et al., 2015; Vidal and Horne, 2003). This study aims to better understand how tubifex worms affect HCHs fluxes from overlying water to the sediment and assess the bioaccumulation and elimination of HCH isomers in Tubifex tubifex. Four experimental treatments were studied. For the bioaccumulation test, three experimental conditions were analyzed: contaminated water column with or without worms and uncontaminated water column with worms. The last one was the poisonous worms eliminated HCHs treatment. This research investigates the effects of tubifex on fluxes of HCHs between the water column and sediment compartments. The different bioaccumulation and elimination behaviors of α-HCH and β-HCH in tubifex were observed in the present study. Moreover, the enantioselectivity of α-HCH was also assessed.

2. Materials and methods 2.1. Chemicals and reagents

α-, β-hexachlorocyclohexane (HCH) pesticide standards (purity 4 99.0%) and recovery surrogate 2,4,5,6-tetrachloro-m-xylene (TCmX) were obtained from J&K Scientific Ltd (Beijing, China). All test chemicals were initially dissolved in n-hexane. Acetone, n-hexane, and petroleum ether (60–90 °C) were HPLC grade and purchased from commercial sources (Beijing Chemical Reagent Co., China). 2.2. Test worms and sediment T. tubifex was obtained from a commercial breeder (Da Senlin Flower Market Beijing, China). Worms were maintained in 2 L plastic tanks containing uncontaminated sediment and dechlorinated tap water at 20 71 °C, with 12 h light and 12 h darkness per day for one week to acclimate to the environment. The worms were fed TetraMin Flakes (Tetra Werke, Melle, Germany) weekly. For the experiments, adult T. tubifex (aged five to seven weeks) were used, and the organisms were not fed during the experimental period. The sediment was collected from Ulla Gail Lake (Inner Mongolia, China) from the 0–10 cm surface layer of sediment. No HCHs were found at detectable levels in the sediment. After freezedrying, sediment samples were homogenized and sieved through a 2 mm sieve, and kept in the dark. The physicochemical properties of the sediment were as follows: clay, 1.81 70.04%; silt, 7.39 72.22%; sand, 90.80 72.26%; organic matter, 5.31 70.21%; and pH, 9.567 0.05. 2.3. Bioaccumulation and elimination experiment Experimental design. The experiments were conducted with four scenarios, which were denoted as þTSE,  TSE, TSE and EE. The experiments were conducted in 500 mL beakers. Sediment

150 g (water content 35%) was put into the bottom of each beaker with a height of 2–3 cm, and 100 mL dechlorinated tap water with HCHs (0.1 mg/L) was added slowly. The scenario with acclimated tubifex worms (10 g) were placed into beakers (18 beakers, 6 sampling points, triplicates for one sampling point). After an exposure period (1, 2, 3, 5, 7, and 10 days), overlying water was gently poured and sampled firstly. The alive worms were collected by heating the beakers in a water bath (40 °C, 30–45 min), which caused them rising to the surface of sediment (Lagauzere et al., 2009; OECD, 2005). Removed the worms from the beakers and rinsed with dechlorinated tap water, then dried the peripheral water using absorbent paper. Three replicates including sediments, overlying water and worms were sampled on the days of 1, 2, 3, 5, 7, and 10 during the 10-d exposure period, weighed and stored at  20 °C. The scenario including contaminated water column with tubifex and uncontaminated sediment was denoted as þ TSE. To investigate the bioturbation of tubifex, a separate experiment (negative control) with spiked water and sediment was denoted as  TSE, and this treatment was without tubifex. The other control scenario, including tubifex, water and sediment without HCHs, was denoted as TSE. Sample modes were consistent with þTSE treatments. The elimination experiment (EE) was conducted after three days of bioaccumulation experimentation; worms were removed to beakers with contaminant-free overlying water and sediment. Sampling was conducted at 0.5, 1, 3, 5, and 7 days. Overlying water, sediment and worms were collected and stored at  20 °C. The treatments had three replicates. For the four scenarios, the test beakers were weighed daily, and the loss of water resulting from evaporation was compensated by adding dechlorinated tap water. All of the beakers were cultured with a randomized block in the climate chamber at 2071 °C and a light-dark cycle of 12 h. Water quality parameters including pH, ionic potential, conductivity, total dissolved solids, salinity and resistivity in the overlying water were measured before sampling, using a SX736 portable pH/conductivity/dissolved oxygen meter (Shanghai San-Xin Instrumentation, Inc.). 2.4. Samples extraction and purification All the samples were thawed at room temperature. The overlying water samples (20 g) were extracted with 20 mL of petroleum ether in a 50 mL polypropylene centrifuge tube. After stirring for 3 min on a vortex mixer, the extraction was repeated twice using fresh solvent. The combined solvent phase was filtered through 5 g of anhydrous sodium sulfate for dehydration, transferred to a pear-shaped flask, and then evaporated to dryness at 35 °C by vacuum rotary evaporator (Shanghai Yarong Biochemistry Instrument Factory, Shanghai, China). The extractive was redissolved with 500 mL n-hexane after passing through a filter membrane (pore size, 0.45 μm). Microwave-assisted solvent extraction (MAE) was applied to extract the tubifex and sediment samples. MAE was carried out with a Mars 6 Xpress (maximum power: 1800 W) microwave extraction system (CEM, Matthews, NC, USA). Tubifex samples were homogenized with an Ultra-Turrax T18 homogenizer for 30 s, and 3 g homogenate was transferred to the microwave extraction vessels (TCmX was added as recovery surrogate), adding 20 mL extraction solution (acetone: petroleum ether ¼ 1:1, V:V). The vessels were covered with pressure-resistant holders and placed into the rotary base. The extraction temperature program was as follows: ramped to 100 °C in 5 min, held for 10 min and cooled to room temperature. The extraction solvent was filtered through 5 g of anhydrous sodium sulfate for dehydration and transferred to a pear-shaped flask, then concentrated to dryness. The Florisil-SPE

S. Di et al. / Ecotoxicology and Environmental Safety 126 (2016) 163–169

cartridge (1000 mg, 6 mL, Agilent SampliQ Products) was used to clean up interfering substances. The cartridge was eluted with 10 mL of acetone, 5 mL of n-hexane, and then equilibrated with 5 mL leachate (acetone: n-hexane ¼ 1:15, v/v). The sample was loaded to the cartridge with 2 mL leachate, and then the cartridge was eluted with 6 ml leachate. The eluate was concentrated to dryness under a gentle nitrogen flow, and diluted with 500 mL of n-hexane. The sediment samples (10 g) were extracted with the same method above.

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tubifex and water samples (SPSS 20.0, IBM, Chicago, USA, po 0.05). The concentrations among the HCH isomers were analyzed using one-way analysis of variance (one-way ANOVA), and the Student–Newman–Keuls (S–N–K) test was used to compare results at p o0.05.

3. Results and discussion 3.1. Bioaccumulation of HCHs in tubifex

2.5. Chemical analysis HCH isomers were analyzed using an Agilent 7890A gas chromatograph equipped with a nickel-63 electron capture detector (μECD) and a HP-5 column (30 m, 0.32 mm i.d., 0.25 μm film; Agilent Technologies Inc.). The flow rate of carrier gas high-purity nitrogen was 1 mL/min. The injector and detector temperatures were 270 °C and 290 °C, respectively. The oven temperature began at 100 °C (hold 2 min), increased to 180 °C at 15 °C/min (hold 5 min), to 185 °C at 5 °C/min (hold 2 min), to 270 °C at 20 °C/min, and finally increased to 290 °C at 20 °C/min (hold 10 min). The enantiomers of α-HCH were analyzed using an Agilent 7890A GC-μECD equipped with a BGB-172 (30 m, 0.25 mm i.d., 0.25 mm film; BGB Analytik AG, Switzerland) chiral column. Nitrogen was applied as carrier gas at a flow rate of 0.7 mL/min. The oven temperature program was as follows: the injector and detector temperatures were 250 °C and 280 °C, respectively. Initial temperature 60 °C, increased to 150 °C at 15 °C/min, to 176 °C at 0.8 °C/min (hold 10 min), to 180 °C at 2 °C/min (hold 4 min), and finally increased to 220 °C at 10 °C/min (hold 1.5 min). (  )-α-HCH eluted first according to Falconer et al. (1997). The enantiomer fraction (EF) was used to measure the enantioselective behavior of α-HCH enantiomers in our experiment. EF was expressed as the proportion of (  )-α-HCH to the sum of (þ)-α-HCH and ( )-αHCH peak areas. The EF values range from 0 to 1, and EF ¼0.502 represents the racemic mixture for α-HCH. The average recoveries for HCHs ranged from 75.3% to 110.0% in tubifex tissue (0.01–0.1 mg/kg), from 92.3% to 109.9% in sediment (0.01–0.1 mg/kg), and from 73.7% to 108.8% in water (0.001– 0.01 mg/kg), with SD below 20% (n ¼3 for each sample type). The limit of detection (LOD, which was defined as the concentration that produced a signal-to-noise ratio of 3) was 0.35 mg/kg and 0.82 mg/kg for α-HCH and β-HCH, respectively.

During a 10-d exposure period, the changes of α-HCH and βHCH concentrations in tubifex tissue showed an “M” trend (Fig. 1A). The first decrease in HCHs concentrations occurred on the 5th day, when some worms might have autotomized. ParisPalacios et al. (2010) found that isoproturon residue was higher in the posterior part of tubifex, which was autotomized, than in the rest of the body. In addition, most worms that had autotomized before the 4th day had the time to regenerate before the 7th day. Lucan-Bouché et al. (1999) also reported that plumbum (Pb) concentrations in the posterior of tubifex were up to 40 times higher than those in the ‘heads’. Furthermore, the worm could lose up to one third of its total length and its capacity for regeneration was not impaired in the contaminated medium. Thus, it can be assumed that, most HCHs were accumulated in the posterior part

2.6. Data analysis The elimination kinetics of HCHs in tubifex samples were simulated using a zero-order kinetics model (Origin8.5) (Galperin and Kaplan, 2008).

C(t ) =C(t = 0)−kt where C(t) (mg/kgwwt) is the concentration of HCHs in the tubifex at sample time t (days), and k is the elimination rate constant of the chemical. The corresponding half-life (t1/2) was calculated by t1/2 ¼C(t ¼ 0)/2k. In this work, we use accumulation factor (AF) to express the relative sorptive capacities of tubifex versus the surrounding environment. The term accumulation factor is used for any time point of uptake when steady state has not been reached (Egeler. et al., 2001). The AF is defined as:

AF=Cworm /Csediment where Cworm and Csediment are the concentration of HCH isomers in tubifex and sediment respectively. A one-sample t test was used to compare the EF values in

Fig. 1. Concentrations of HCHs in the worms (A), overlying water (B) and sediment (C) for the þ TSE and  TSE treatments. Bars are standard error.

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Table 2 The EF values of α-HCH enantiomers in tubifex, overlying water and sediment for the þ TSE,  TSE and EE treatments. Enantiomer factor EF

Exposure time (days) 1

2

3

5

7

10

Worms-( þ TSE) Overlying water-( þTSE) Sediment-( þTSE) Overlying water-(  TSE) Sediment-(  TSE)

0.4907 0.001 0.5277 0.012 0.5077 0.003 0.503 70.007 0.509 70.006

0.4917 0.001 0.552 7 0.009 0.4847 0.002 0.5317 0.001 0.4867 0.002

0.496 70.003 0.5317 0.011 0.482 70.003 0.5127 0.004 0.4797 0.002

0.5017 0.002 0.529 7 0.010 0.492 70.002 0.5247 0.001 0.494 70.003

0.4887 0.001 0.505 70.004 0.4887 0.001 0.5077 0.001 0.4897 0.003

0.4987 0.001 0.540 7 0.006 0.4897 0.003 0.4997 0.005 0.4947 0.002

Enantiomer factor EF

Exposure time (days) 0.5 0.502 70.001 0.492 7 0.005 0.492 7 0.001

1 0.5007 0.002 0.494 7 0.003 0.4997 0.005

3 0.503 7 0.001 0.494 70.007 0.501 70.001

5 0.5017 0.003 0.502 7 0.001 0.503 7 0.004

7 0.503 70.001 0.502 70.006 0.5047 0.004

Worms-(EE) Overlying water-(EE) Sediment-(EE)

would reduce the bioavailability of α-HCH. Thus, the differences were mainly caused by different uptake rate constants, as reported by Rinderhagen and Butte (1995). The accumulation factor (AF) values of the two isomers were shown in Table 1. The AFs of β-HCH in tubifex were higher than those of α-HCH, which was in accordance with the report of Rinderhagen and Butte (1995). The differences in the orientation of the chloride and hydrogen atoms at the cyclohexane ring lead to different stereoisomers with different physicochemical properties (such as volatilities and partition coefficients). Thus, the differences in AFs may be caused by different steric configurations, which affected adsorption to the integument and to permeation of cell membranes (Rinderhagen and Butte, 1995). The other speculative reason may be the stereoisomer isomerization of HCHs, leading to higher AFs for the most stable isomers. For example, Wu et al. (1997) found that α-HCH could be converted to β-HCH in the environment. However, the AFs in this study were lower than those in water-only treatments, which were 2530 7 665 and 14,630 71465 for α-HCH and β-HCH respectively (Rinderhagen and Butte, 1995). These phenomena may have resulted from the hydrophobicity and low polarity of HCHs, HCHs tend to be associated with soils and sediments once released in the environment (Duan et al., 2008). Thus, most HCHs were combined with sediment in þTSE treatment, whereas most HCHs were adsorbed by tubifex in water-only treatment. Egeler et al. (1997) reported that the AFs of the resulting worm were relatively low due to high sediment concentrations. Another reason may be the existence of detoxification (autotomy and regeneration) in the þTSE treatment. Tubifex in the sediment might autotomize and excrete most HCHs. Furthermore, the changes of feeding physiology of tubifex to adapt to the low food availability in the water-only test, may lead to higher assimilation efficiencies of HCHs as carbon sources. Hence, the relative bioaccumulation capacities of HCHs in tubifex are related to different environments. Among the HCH isomers, α-HCH is the only one with chirality, and it has two enantiomers. The EF values of α-HCH enantiomers in tubifex were calculated, and the data were shown in Table 2. A one-sample t-test was carried out to compare the EF values in tubifex with EF ¼0.502, and there was no significant deviation

of tubifex, and worms had autotomized before the 5th day, which led to the first decrease. Then worms had regenerated before the 7th day and accumulated more HCHs. The second increase in HCHs concentrations on the 7th day (0.34 mg/kgwwt for α-HCH and 0.87 mg/kgwwt for β-HCH, respectively) was higher than the first increase (on the 3 rd day, 0.27 mg/kgwwt for α-HCH and 0.63 mg/ kgwwt for β-HCH). Vidal and Horne (2003) reported that tubifex exposed to the high levels of mercury (Hg) in sediment had high mercury tolerance, and this tolerance was genetically persistent. Thus, it can be assumed that, the posterior part after regeneration had higher HCHs tolerance, which led to the accumulation of more HCHs. The second decrease may be due to both autotomy and the worm’s increase in inactivated efficiency. Moreover, LuszczekTrojnar et al. (2014) reported cadmium (Cd) induced mucus secretion increasing in tubifex, and it was considered as an adaptive response related to the physiological resistance phase. When emitted at the surface of the body mucus containing the contaminants represents a detoxification mechanism. However, mucus produced in large amounts may induce physiological disturbances in gas exchange, which may lead to death. Thus, the inactivated efficiency may be caused by the toxic effects of HCHs and the increase of mucus. However, the model of bioaccumulation curves in tubifex were different from the model of epoxiconazole (Liu et al., 2014a) and triadimefon (Liu et al., 2014b), in which a simple one-peak shaped accumulation curve was observed in the spiked soil test. Different models of bioaccumulation curves may represent different storage, detoxification, metabolism and excretion patterns in tubifex when this organism encountered different contaminations (Liu et al., 2015). Moreover, the differences may be attributed to repetition at different concentrations (Di et al., 2014), and small differences in the test conditions may significantly influence the results (Luszczek-Trojnar et al., 2014). Furthermore, the accumulation concentrations of β-HCH were significantly higher than those of αHCH. The reason may be that β-HCH has the lowest vapor pressure among the HCH isomers, and tends to accumulate in animal tissue (Syed et al., 2013). Working α-HCH along with β-HCH in dechlorinated tap water may cause interactions with each other or other substances, and they may compete for biological sites, which Table 1 The accumulation factor of HCH isomers in tubifex for þTSE treatments. Accumulation factor AF

α-HCH-(þ TSE) β-HCH-(þ TSE)

Exposure time (days) 1

2

3

5

7

10

3.7 7 0.3 6.17 0.7

2.5 70.1 4.2 70.3

12.7 71.1 23.3 71.7

11.5 70.6 11.7 70.1

16.4 7 0.2 30.17 1.0

9.9 7 0.1 13.6 7 0.4

S. Di et al. / Ecotoxicology and Environmental Safety 126 (2016) 163–169

from 0.502. Therefore, the bioaccumulation of α-HCH enantiomers in tubifex tissue was not enantioselective.

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Table 3 The kinetic parameters of zero-order kinetics equation for elimination processes in tubifex.

3.2. Biological effect of tubifex

C(t) ¼ C(t ¼ 0)-kt

Several authors have reported that the existence of tubifex may be an important factor influencing the dissipation of pesticides in sediment, and tubifex played an important role in refining the contaminated sediment (Ciutat et al., 2005a; Ciutat et al., 2005b; Liu et al., 2015). Ciutat et al. (2005a) found that the bioturbation of tubifex increases the cadmium (Cd) fluxes from overlying water to freshwater sediment. Thus, the existence of tubifex may be an important factor influencing the distribution of HCHs in the overlying water and sediment. There were significant differences in the concentrations of β-HCH between worm-present treatments (þ TSE) and control treatments (  TSE) (Fig. 1B and C). Tubifex decreased β-HCH fluxes from overlying water into sediment. The concentrations of β-HCH in overlying water and sediment in the -TSE treatments were 5.8 and 1.4 times higher than those in the þTSE treatments. The lower concentrations may be due to the accumulation and degradation by worms. The removal values of βHCH in overlying water over a 24-h period for the þTSE and  TSE treatments were approximately 89% and 40%, respectively. The decrease for the  TSE treatments in the 1st day may be attributed to photodecomposition, evaporation and adsorption onto sediment. The adsorption of β-HCH on sediment was 36.6% and 33.8% for  TSE and þ TSE treatments, respectively. The additional losses in þTSE treatments may be due to some factors such as uptake or degradation by tubifex. To determine whether β-HCH is: (a) absorbed and stored in tubifex, or (b) degraded by tubifex, residual concentrations were measured in tubifex. The accumulation of β-HCH in tubifex was 13.7%, so most β-HCH loss in the overlying water was due to metabolism by tubifex. This result was consistent with the quick elimination of β-HCH in worms. Furthermore, the existence of tubifex did not have significant effects on α-HCH fluxes, and the concentrations of α-HCH were all lower than those of β-HCH for these treatments (þTSE and TSE), which may be explained by the different physicochemical properties. Among the HCH isomers, α-HCH is more likely to partition to the air and to transport over long distances due to its higher volatility and lower partition coefficient, whereas β-HCH is more resistant to hydrolysis and has the lowest vapor pressure among all of the HCH isomers (Syed et al., 2013). Moreover, α-HCH can be transformed into β-HCH in the natural environment (Wu et al., 1997). Significant enantioselectivity of α-HCH was found in the overlying water for þ TSE, and the EF values were shown in Table 2. The result indicated that the (þ )-α-HCH was preferentially eliminated over the ( )-α-HCH in water. The EFs in worms and sediment (þ TSE) were slightly below 0.502, which were just the opposite of those in the overlying water. Thus, tubifex and sediment may be the reason for the appearance of enantioselectivity. We speculated that tubifex excreted α-HCH enantiomers into the water enantioselectively and that the microorganisms in sediment may enantioselectively degrade α-HCH enantiomers, but these possibilities were not verified in the present work and need further study.

C(t ¼ 0)

k

R2

t1/2

0.4407 0.010 0.582 7 0.001

0.04977 0.0018 0.0540 7 0.0044

0.993 0.967

4.43 5.39

3.3. Elimination of HCHs in tubifex After three days of accumulation, the depuration experiment commenced. Elimination kinetics followed a zero-order kinetics model well (R2 ¼0.967–0.993), and the kinetics parameters were shown in Table 3. As illustrated in Fig. 2A, approximately 80% and 70% of α-HCH and β-HCH were depleted or excreted after seven days, respectively. The concentrations of β-HCH were significant

α-HCH-(EE) β-HCH-(EE)

Fig. 2. Elimination curves for HCH isomers in tubifex tissue (A), overlying water (B) and sediment (C), bars are standard error.

higher than those of α-HCH during the 7-d elimination experiment, which may be due to the different bioaccumulation concentrations (0.440 and 0.582 mg/kgwwt for α-HCH and β-HCH respectively) in tubifex on the 3 rd day. The α-HCH residue was significant decreased from the beginning of EE, while the β-HCH residue was significant decreased after 0.5 day (ANOVA, S–N–K, po 0.05). The elimination rate constants for α-HCH and β-HCH were calculated and the depuration half-lives were 4.43 and 5.39 days respectively, confirming their rapid elimination from tubifex. The rapid elimination may be explained by biotransformation, excretion or autotomy. Paris-Palacios et al. (2010) reported that the induction of autotomy was not stopped by placing the worms in

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Table 4 The water quality parameters of overlying water in þ TSE,  TSE, TSE and EE treatments. Treatments

Days units

pH

þTSE

1 2 3 5 7 10

8.15 7.90 7.80 7.76 8.10 8.20

 TSE

1 2 3 5 7 10

TSE

1 2 3 5 7 10

EE

0.5 1 3 5 7

mV mV

COND mS/cm

TDS g/L

SALL ng/L(ppt)

RES MΩ cm

 80  66  60  58  78  83

1.94 3.28 3.28 4.25 5.32 4.30

1.42 2.49 2.50 3.23 4.07 3.29

0.99 1.70 1.70 2.22 2.81 2.26

509 303 303 236 188 231

8.49 8.46 8.59 8.58 8.89 8.85

 99  98  105  106  123  119

1.92 2.51 2.16 2.25 2.87 2.71

1.39 1.86 1.58 1.65 1.98 2.01

0.97 1.28 1.09 1.14 1.36 1.38

521 398 462 444 375 371

8.08 7.88 7.58 7.64 8.19 8.42

 75  64  46  50  82  95

2.95 3.62 4.46 6.25 6.52 6.40

2.22 2.77 3.39 4.92 5.18 5.09

1.52 1.88 2.33 3.33 3.46 3.41

338 276 224 159 154 155

8.46 8.47 8.33 8.41 8.84

 92  92  85  89  113

3.64 3.71 4.15 5.30 5.67

2.79 2.85 3.17 4.06 4.38

1.89 1.93 2.17 2.80 3.00

274 268 240 188 176

mV: ionic potential; COND: conductivity; TDS: total dissolved solids; SALL: salinity; RES: resistivity.

clean water, and regeneration took place simultaneously with an important decontamination pattern of tubifex. The total amount of HCHs released into the water and sediment during the seven days were less than 15.1% (α-HCH) and 14.8% (β-HCH) of the originally accumulated HCHs in the tubifex tissue. Thus, if the evaporation was not considered, the majority of the compounds were metabolized by the organism. The results indicate that high HCHs concentrations accumulated in worms after the exposure dissipated during rapid and effective purification. This means that contaminated worms after purification may be safe for the subsequent organisms of the food chain. The concentrations of HCHs in overlying water had changed little in three days after being excreted by tubifex, then the concentration increased to 0.0014 and 0.0019 mg/L for α-HCH and βHCH respectively (Fig. 2B). The concentrations of HCHs in sediments exhibited time dependent increases (Fig. 2C), which were in accordance with the linear decrease in HCHs concentrations in tubifex. Furthermore, the concentrations of β-HCH in overlying water and sediment were all higher than those of α-HCH, which were in accordance with the concentrations in tubifex (Fig. 2A). Additionally, enantioselectivity of α-HCH was not found for all of the elimination experiments (Table 2). 3.4. Water characterization Water quality parameters including pH, ionic potential and resistivity were significantly different among worm-present (þ TSE and TSE) and worm-free (  TSE) treatments (Table 4). Resistivity, pH and the absolute values of ionic potential in  TSE treatments were higher than those in the þTSE and TSE treatments. Furthermore, pH and the absolute values of ionic potential showed the same tendency in the worm-present treatments (þ TSE and TSE), which could be described as a “decrease-

increase” process, whereas the tendency was increasing in the worm-free (  TSE) treatments. The conductivity, total dissolved solids and salinity followed the same trend in the þ TSE, TSE,  TSE and EE treatments, which increased with exposure time. Moreover, the resistivity of all treatments were decreasing with exposure time. Compared with other treatments, no significant changes in pH and ionic potential were observed for the elimination experiment. Thus, the existence of tubifex may have effects on the pH and ionic potential of water. Kaonga. et al. (2010) reported that the range of pH in tubifex was 7.3–8.9; thus, it can be speculated that, the pH and ionic potential of water were connected with the bodily fluid of the worms. The relationship between the concentrations of HCHs in tubifex and water properties were investigated. Significant correlations were found between β-HCH concentrations in water and conductivity (r ¼  0.945, p ¼0.004), total dissolved solids (r ¼  0.945, p¼ 0.004), salinity (r ¼ 0.943, p ¼0.005), and resistivity (r ¼0.939, p¼ 0.006), but not between β-HCH concentrations and pH (r ¼  0.109, p¼ 0.837), ionic potential (r ¼0.126, p ¼0.812), and αHCH concentrations (r ¼0.688, p ¼0.131) in þTSE treatments. Significant correlations between β-HCH concentrations and water properties (r o  0.825 or r 40.846, p o0.043) were also found in  TSE treatments. Thus the α-HCH and β-HCH concentrations were significantly correlated with conductivity, total dissolved solids and salinity, which may be influenced by the physicochemical properties of HCHs. HCHs have a strong affinity for suspended particulates and adsorption sites in sediments. The changes of water quality parameters may impact the bioavailability of nutrients and other pollutions, thus impacting the aquatic environment.

4. Conclusion In this study, the bioaccumulation and elimination of α-HCH and β-HCH in tubifex from spike-water with sediment were investigated. The M-type concentration curves were found in tubifex, and the depuration half-lives were 4.43 (α-HCH) and 5.39 (βHCH) days respectively. The results indicated that worms might protect themselves against the increase in internal concentrations of HCHs by autotomizing the caudal region in which the HCHs were mostly accumulated. The capacity of regeneration might be not impaired, even in contaminated media. Moreover, tubifex rapidly reduced HCHs concentrations, which suggests that they are a safe food for subsequent organisms. The tubifex had a higher capacity to accumulate β-HCH than α-HCH, and significantly reduced β-HCH fluxes from overlying water to sediment. Worms may have effects on the pH and ionic potential in the overlying water, and HCHs concentrations may influence the conductivity, total dissolved solids and salinity of water.

Acknowledgments This work was supported by funding from the National Natural Science Foundation of China (Contract Grant nos: 21177154 and 41201499).

Appendix A. Supplementary material Supplementary data associated with this article can be found in the online version at http://dx.doi.org/10.1016/j.ecoenv.2015.12. 022.

S. Di et al. / Ecotoxicology and Environmental Safety 126 (2016) 163–169

References Ali, I., Abdul-Enein, Hassan Y., 2004. Chiral Pollutants: Distribution, Toxicity and Analysis by Chromatography and Capillary Electrophoresis, Proteomics. Ali, I., Jain, C.K., 1998. Groundwater contamination and health hazards by some of the most commonly used pesticides. Curr. Sci. 75, 1011–1014. Chessells, M.J., Hawker, D.W., Connell, D.W., Papajcsik, I.A., 1988. Factors influencing the distribution of lindane and isomers in soil of an agricultural environment. Chemosphere 17, 1741–1749. Ciutat, A., Anschutz, P., Gerino, M., Boudou, A., 2005a. Effects of bioturbation on cadmium transfer and distribution into freshwater sediments. Environ. Toxicol. Chem. 24, 1048–1058. Ciutat, A., Gerino, M., Mesmer-Dudons, N., Anschutz, P., Boudou, A., 2005b. Cadmium bioaccumulation in Tubificidae from the overlying water source and effects on bioturbation. Ecotoxicol. Environ. Saf. 60, 237–246. Di, S., Liu, T., Lu, Y., Zhou, Z., Diao, J., 2014. Enantioselective bioaccumulation and dissipation of soil-associated metalaxyl enantiomers in tubifex. Chirality 26, 33–38. Duan, L., Zhang, N., Wang, Y., Zhang, C., Zhu, L., Chen, W., 2008. Release of hexachlorocyclohexanes from historically and freshly contaminated soils in China: implications for fate and regulation. Environ. Pollut. 156, 753–759. Egeler, P., Römbke, J., Meller, M., Knacker, T., Franke, C., Studinger, G., Nagel, R., 1997. Bioaccumulation of lindane and hexachlorobenzene by tubificid sludgeworms (Oligochaeta) under standardised laboratory conditions. Chemosphere 35, 835–852. Egeler, P., Meller, M., Roembke, J., Spoerlein, P., Streit, B., Nagel, R., 2001. Tubifex tubifex as a link in food chain transfer of hexachlorobenzene from contaminated sediment to fisht. Hydrobiologia 463, 171–184. Falconer, R.L., Bidleman, T.F., Szeto, S.Y., 1997. Chiral pesticides in soils of the Fraser Valley, British Columbia. J. Agric. Food Chem. 45, 1946–1951. Fangtong, W., Juan, G., Shiquan, S., Wanchun, T., Shujuan, W., 2011. The Influence of Tubifex bioturbation on total nitrogen release from sediment of Dongting Lake. In: Proceedings of the International Conference on Computer Distributed Control and Intelligent Environmental Monitoring (CDCIEM), IEEE. pp. 2314– 2317. Galperin, Y., Kaplan, I.R., 2008. Zero‐order kinetics model for the Christensen– Larsen method for fugitive fuel age estimates. Ground Water Monit. Remediat. 28, 94–97. Hussain, I., Alothman, Z.A., Alwarthan, A.A., Sanagi, M.M., Ali, I., 2015. Chiral xenobiotics bioaccumulations and environmental health prospectives. Environ. Monit. Assess. 187, 1–23. Jones, K.C., De Voogt, P., 1999. Persistent organic pollutants (POPs): state of the science. Environ. Pollut. 100, 209–221. Kaonga, C.C., Kumwenda, J., Mapoma, H.T., 2010. Accumulation of lead, cadmium, manganese, copper and zinc by sludge worms; Tubifex tubifex in sewage sludge. Int. J. Environ. Sci. Technol. 7, 119–126. Lagauzere, S., Terrail, R., Bonzom, J.M., 2009. Ecotoxicity of uranium to Tubifex tubifex worms (Annelida, Clitellata, Tubificidae) exposed to contaminated sediment. Ecotoxicol. Environ. Saf. 72, 527–537. Liu, A.-X., Lang, Y.-H., Xue, L.-D., Liao, S.-L., Zhou, H., 2009. Probabilistic ecological risk assessment and source apportionment of polycyclic aromatic hydrocarbons in surface sediments from Yellow Sea. Bull. Environ. Contam. Toxicol. 83, 681–687. Liu, C., Wang, B., Xu, P., Liu, T., Di, S., Diao, J., 2014a. Enantioselective Determination of Triazole Fungicide Epoxiconazole Bioaccumulation in Tubifex Based on HPLC-MS/MS. J. Agric. Food Chem. 62, 360–367. Liu, T., Diao, J., Di, S., Zhou, Z., 2014b. Stereoselective bioaccumulation and metabolite formation of triadimefon in Tubifex tubifex. Environ. Sci. Technol. 48, 6687–6693. Liu, T., Diao, J., Di, S., Zhou, Z., 2015. Bioaccumulation of isocarbophos enantiomers from laboratory-contaminated aquatic environment by tubificid worms. Chemosphere 124, 77–82. Lu, X., Chen, C., Zhang, S., Hou, Z., Yang, J., 2013. Concentration levels and ecological risks of persistent organic pollutants in the surface sediments of Tianjin coastal

169

area, China. Sci. World J. 2013, 417435. Lucan-Bouché, M.-L., Biagianti-Risbourg, S., Arsac, F., Vernet, G., 1999. An original decontamination process developed by the aquatic oligochaete Tubifex tubifex exposed to copper and lead. Aquat. Toxicol. 45, 9–17. Luszczek-Trojnar, E., Sroka, K., Klaczak, A., Nowak, M., Popek, W., 2014. Bioaccumulation and purification of cadmium in Tubifex tubifex. Turk. J. Fish. Aquat. Sci. 14, 939–946. Meller, M., Egeler, P., mbke, J.R., Schallnass, H., Nagel, R., Streit, B., 1998. Short-term toxicity of lindane, hexachlorobenzene, and copper sulfate to tubificid sludgeworms(Oligochaeta) in artificial media. Ecotoxicolgy Environ. Saf. 39, 10–20. Nakata, H., Kawazoe, M., Arizono, K., Abe, S., Kitano, T., Shimada, H., Li, W., Ding, X., 2002. Organochlorine pesticides and polychlorinated biphenyl residues in foodstuffs and human tissues from China: status of contamination, historical trend, and human dietary exposure. Arch. Environ. Contam. Toxicol. 43, 0473–0480. Niu, L., Xu, C., Yao, Y., Liu, K., Yang, F., Tang, M., Liu, W., 2013. Status, Influences and Risk Assessment of Hexachlorocyclohexanes in Agricultural Soils Across China. Environ. Sci. Technol. 47, 12140–12147. OECD, 2005. (Organisation for Economic Cooperation and Development). Validation of a Sediment Toxicity Test With the Endobenthic Aquatic Oligochaete Lumbriculus Variegatus by an International Ring Test. Paris-Palacios, S., Mosleh, Y.Y., Almohamad, M., Delahaut, L., Conrad, A., Arnoult, F., Biagianti-Risbourg, S., 2010. Toxic effects and bioaccumulation of the herbicide isoproturon in Tubifex tubifex (Oligocheate, Tubificidae): a study of significance of autotomy and its utility as a biomarker. Aquat. Toxicol. 98, 8–14. Reynoldson, T.B., Rodriguez, P., Madrid, M.M., 1996. A comparison of reproduction, growth and acute toxicity in two populations of Tubifex tubifex (Müller, 1774) from the North American Great Lakes and Northern Spain. Hydrobiologia. 334, 199–206. Rinderhagen, M., Butte, W., 1995. Kinetics of accumulation and elimination of isomeric hexachlorocyclohexanes by tubificids. SAR QSAR Environ. Res. 4, 131–138. Syed, J.H., Malik, R.N., Liu, D., Xu, Y., Wang, Y., Li, J., Zhang, G., Jones, K.C., 2013. Organochlorine pesticides in air and soil and estimated air–soil exchange in Punjab, Pakistan. Sci. Total Environ. 444, 491–497. UNEP, Stockholm Convention on Persistent Organic Pollutant (POPs). 2010. 〈http:// chm.pops.int/Convention/ThePOPs/TheNewPOPs/tabid/2511/Default.aspx〉. Vidal, D.E., Horne, A.J., 2003. Inheritance of mercury tolerance in the aquatic oligochaete Tubifex tubifex. Environ. Toxicol. Chem. 22, 2130–2135. Wu, C., Zhang, A., Liu, W., 2013. Risks from sediments contaminated with organochlorine pesticides in Hangzhou, China. Chemosphere 90, 2341–2346. Wu, W.Z., Xu, Y., Schramm, K.W., Kettrup, A., 1997. Study of sorption, biodegradation and isomerization of HCH in stimulated sediment/water system. Chemosphere 35, 1887–1894. Yang, Y.H., Zhou, S.S., Li, Y.Y., Xue, B., Liu, T., 2011. Residues and chiral signatures of organochlorine pesticides in sediments from Xiangshan Bay, East China Sea. J. Environ. Sci. Health B 46, 105–111. Zeng, E.Y., Venkatesan, M., 1999. Dispersion of sediment DDTs in the coastal ocean off southern California. Sci. Total Environ. 229, 195–208. Zhang, A., Liu, W., Yuan, H., Zhou, S., Su, Y., Li, Y.F., 2011. Spatial distribution of hexachlorocyclohexanes in agricultural soils in Zhejiang province, China, and correlations with elevation and temperature. Environ. Sci. Technol. 45, 6303–6308. Zhang, F., He, J., Yao, Y., Hou, D., Jiang, C., Zhang, X., Di, C., Otgonbayar, K., 2013. Spatial and seasonal variations of pesticide contamination in agricultural soils and crops sample from an intensive horticulture area of Hohhot, North-West China. Environ. Monit. Assess. 185, 6893–6908. Zheng, X., Liu, X., Liu, W., Jiang, G., Yang, R., 2009. Concentrations and source identification of organochlorine pesticides (OCPs) in soils from Wolong Natural Reserve. Chin. Sci. Bull. 54, 743–751. Zhou, Q., Wang, J., Meng, B., Cheng, J., Lin, G., Chen, J., Zheng, D., Yu, Y., 2013. Distribution and sources of organochlorine pesticides in agricultural soils from central China. Ecotoxicol. Environ. Saf. 93, 163–170.