Uptake kinetics of pesticides chlorpyrifos and tebuconazole in the earthworm Eisenia andrei in two different soils

Uptake kinetics of pesticides chlorpyrifos and tebuconazole in the earthworm Eisenia andrei in two different soils

Environmental Pollution 236 (2018) 257e264 Contents lists available at ScienceDirect Environmental Pollution journal homepage: www.elsevier.com/loca...

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Environmental Pollution 236 (2018) 257e264

Contents lists available at ScienceDirect

Environmental Pollution journal homepage: www.elsevier.com/locate/envpol

Uptake kinetics of pesticides chlorpyrifos and tebuconazole in the earthworm Eisenia andrei in two different soils*  ta Svobodova , Kla ra Smídov , Martina Hve zdova , Jakub Hofman* Marke a Research Centre for Toxic Compounds in the Environment (RECETOX), Faculty of Science, Masaryk University, Kamenice 753/5, Brno, CZ-62500, Czech Republic

a r t i c l e i n f o

a b s t r a c t

Article history: Received 26 October 2017 Received in revised form 22 January 2018 Accepted 23 January 2018

Agriculture is today indispensably connected with enormous use of pesticides. Despite tough regulation, their entrance into soil cannot be excluded and they might enter soil organisms and plants and continue further to terrestrial food chains. This study was conducted to investigate the bioaccumulation of two pesticides currently used in large amounts, the insecticide chlorpyrifos (CLP) and the fungicide tebuconazole (TBZ). Their detailed uptake kinetics in the model earthworm species Eisenia andrei were measured in two arable soils differing in organic carbon content (1.02 and 1.93% respectively). According to our results, a steady state was reached after 3e5 days for both pesticides and soils. The values of bioaccumulation factors calculated at the steady state ranged from 4.5 to 6.3 for CLP and 2.2e13.1 for TBZ. Bioaccumulation factors were also calculated as the ratio of uptake and elimination rate constants with results comparable with steady-state bioaccumulation factors. The results suggested that the degradation and bioaccumulation of tested compounds might be influenced by other factors than only total organic carbon (e.g. clay content). The lower Koc and hydrophobicity of TBZ relative to CLP probably led to higher availability of TBZ through pore water exposure. On the other hand, CLP's higher hydrophobicity probably caused an increase in availability by its additional uptake via ingestion. To enable a proper ecological risk assessment of current pesticides in soils, it is necessary to accurately determine their bioaccumulation in soil invertebrates. We believe that our study not only brings such information for two specific pesticides but also addresses key methodological issues in this area. © 2018 Elsevier Ltd. All rights reserved.

Keywords: Currently used pesticides Bioaccumulation Toxicokinetics Eisenia andrei Soil Chlorpyrifos Tebuconazole

1. Introduction The large group of currently used pesticides (CUPs), which includes organophosphate insecticides and conazole fungicides for example, has been used since environmentally problematic organochlorinated pesticides (OCPs) were banned because of their high persistence and toxicity. CUPs are supposed to have lower persistence and environmental mobility than OCPs and due to modern pesticide regulation and legislation they should have only minor environmental effects, particularly on local ecosystems. However, these expectations are not always met: many CUPs show half-lives of months (e.g. pendimethalin, boscalid, quinoxyfen) and even years (e.g. epoxiconazole, flusilazole, diflufenican) (PPDB, 2016) and their repeated and massive use in agriculture can lead to gradual


This paper has been recommended for acceptance by Dr. Jorg Rinklebe. * Corresponding author. E-mail address: [email protected] (J. Hofman).

https://doi.org/10.1016/j.envpol.2018.01.082 0269-7491/© 2018 Elsevier Ltd. All rights reserved.

accumulation in soils (pseudo-persistence). Monitoring of arable soils in central Europe revealed that 81% of monitored soils had at least 1 pesticide above 0.01 mg kg1 several months after their last zdova  et al., 2018). Soil can act as a sink for possible application (Hve further distribution of CUPs to other environmental compartments by runoff to surface water, leaching to groundwater or by transfer and possible accumulation in plants and soil biota, resulting in distribution of the toxicant into the terrestrial food chain. Within regulatory risk assessment of pesticides in the EU, there is presently a need for more research on uptake kinetics in in-soil organisms as highlighted by the recent “EFSA Scientific Opinion addressing the state of the science on risk assessment of plant protection products for in-soil organisms” (EFSA, 2017). In this panel, bioaccumulation in soil organisms (earthworms) was considered in order to assess the potential for secondary poisoning in birds and mammals (EFSA, 2009). Earthworms may constitute up to 80% of the total biomass of the soil fauna (Kabata-Pendias, 2010) and they live in close contact with the soil, have a permeable cuticle, and consume large amounts of


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soil. Due to these characteristics they are an appropriate model organism for bioaccumulation assays (Cortet et al., 1999; Jager et al., 2005; OECD, 2010). The majority of bioaccumulation studies with earthworms are focused on heavy metals and persistent organic pollutants. Only a few studies have focused on CUPs and they are limited only to several representative compounds while there are several hundreds of approved CUPs. The findings of the studies indicate that soil organic matter (SOM) and clay contents in soils (Papini et al., 2006; Wu et al., 2011a; Yu et al., 2006), hydrophobicity (Kow) (Chang et al., 2016; Xu et al., 2014; Yu et al., 2006) and chirality (Chen et al., 2014; Diao et al., 2011; Yu et al., 2012) of the compounds are major factors in CUPs bioaccumulation in earthworms. To our knowledge, there is only one study on uptake of tebuconazole (TBZ) by earthworms (Yu et al., 2012), but it focused on enantioselectivity of TBZ bioaccumulation and it was performed only in one soil. Chlorpyrifos (CLP) bioaccumulation in earthworms has been addressed only in four studies (Yu et al., 2006; Wu et al., 2011a; Lister et al., 2011; Spurgeon et al., 2011), but none of them compared the uptake kinetics of CLP in different soils. The two main approaches for assessing bioaccumulation process in earthworms are: i) studies with a fixed time of exposure and determination of the bioconcentration factor (BCF), bioaccumulation factor (BAF) or biota-soil accumulation factor (BSAF) (Gevao et al., 2001; Belden et al., 2004; Yu et al., 2006; Wu et al., 2011a) or ii) studies focused on whole uptake kinetics of CUPs by earthworms with calculation of kinetic parameters e uptake (assimilation) and elimination rate constants (Lister et al., 2011; Spurgeon et al., 2011; Yu et al., 2012). For calculation of BCF, BAF or BSAF, it is assumed that an equilibrium between the earthworm body and soil is reached during the exposure as a result of general equilibrium partitioning (EqP) between all phases present (soil solids, pore water, biota). However, the time needed to reach a steady state can differ between species, chemicals or soils. For some compounds and soils, the time necessary to reach such equilibrium is very long e.g. over 28 days for brominated organic compounds (Nyholm et al., 2010). Also, this approach assumes that the uptake curve follows the first-order kinetics and it is invalid for e.g. peak shape curves related to fast degradation of the compound in soil. In those cases, it is not clear whether the steady state was really reached or not, mostly in studies with PAHs (e.g. Jager et al., 2000;   and Hofman, 2014) but also in CUP Matscheko et al., 2002; Smídov a studies (Jantunen et al., 2008; Yu et al., 2012; Xu et al., 2014). Therefore, Jager et al. (2000) suggested that BCF should be expressed dynamically as a ratio of uptake and elimination rate constants. As far as we know, this approach has been used in studies of CUPs with the aquatic oligochaete Lumbriculus varie€a € et al., 2003; Jantunen et al., 2008) but also the gatus (M€ aenpa earthworm Eisenia fetida (Yu et al., 2012). Moreover, kinetic data for CUPs are rare with more than one tested soil. In our experiment, we studied the uptake kinetics of two CUPs (the insecticide chlorpyrifos and the fungicide tebuconazole) in earthworms Eisenia andrei in two arable soils of different properties. The aims were to find a sufficient length of exposure to achieve equilibrium in concentration between soil and earthworm, to model the uptake kinetics and to compare two pesticides and two soils. TBZ and CLP were selected as model CUPs because of their very high usage in the European Union. In the Czech Republic, the representative Central Europe country, CLP and TBZ are the third and fifth most used synthetic pesticides respectively, with an average consumption in 2012e2014 of 182 t y1 for CLP and 167 t y1 for TBZ. The two compounds have distinct properties (Table 1) suggesting their different fate in soil and bioaccumulation. Two contrasting soils (one containing about twice as much total organic carbon (TOC) and clay than the other) were used to see the possible influence of different soil properties on the uptake of

compounds by earthworms. We also compared our experimentally obtained BAFs with the BCF model acquired according to the EFSA Guidance document “Risk assessment for birds and mammals” (EFSA, 2009) used to assess pesticide risk assessment at the EU level. 2. Materials and methods 2.1. Experimental soils and test organisms Two non-contaminated arable soils of fluvisol type, FS1 and FS2, with different TOCs (1.02% and 1.93%) were sampled in the Czech Republic in the summer of 2011. The properties of the studied soils are summarized in Table 2. Soils were air-dried, sieved to a maximum 2 mm fraction and stored in plastic bags in the dark. Both soils were sterilized by gamma irradiation at 25 kGy before the experiment (Bioster Ltd., Czech Republic) to represent the situation of a higher bioavailable fraction caused by limited microbial degradation in order to simulate the worst-case scenario in the soil   and Hofman, 2014). Sterilized soil was also environment (Smídov a used in the studies of Yu et al. (2006), Wu et al. (2011a, 2011b). Distilled water was added to reach 50% of the water-holding capacity (WHC) of the individual soils 2 days before contamination. Moist soils were stored in airtight glass vessels in the dark at 20 ± 2  C. The earthworms (Eisenia andrei) were from our laboratory (Research Centre for Toxic Compounds in the Environment, Czech Republic). The substrate for the culture was a mixture of granulated manure (40%), garden substrate (50%), and peat (10%) and was moistened to 60e80% WHC. The pH ranged from 6 to 7 as adjusted by CaCO3. Before the experiment, adult earthworms with developed clitellum and a minimal weight of 250 mg were acclimatized in uncontaminated soils moistened to 50% WHC with added manure 7 mg g1 dry weight for 3 days. 2.2. Test chemicals and contamination of soils Chlorpyrifos [O,O-diethyl O-(3,5,6-trichloro-2-pyridyl)phosphorothioate; CAS number 2921-88-2] and tebuconazole [(RS)-1-pchlorophenyl-4,4-dimethyl-3-(1H-1,2,4-triazol-1-ylmethyl)pentan-3-ol; CAS n. 107534-96-3] were obtained as neat standards from Sigma-Aldrich (Germany) with a purity of >99%. A spiking solution of CLP and TBZ in acetone was prepared to achieve a nominal concentration in soils of 5 mg kg1 soil dry weight (kgsoil-dw). Similar concentrations (units to tens of mg.kg1) have frequently been used in other bioaccumulation studies of pesticides (e.g. Yu et al., 2012; Yu et al., 2006; Spurgeon et al., 2011; Wu et al., 2011a, 2011b; Xu et al., 2014; Papini et al., 2006; Belden et al., 2004). When compared to long-term concentrations of peszdova  et al., 2018), the mg.kg1 ticides in agricultural soils (Hve levels used in laboratory bioaccumulation studies are 2-3 orders of magnitude higher. However, they might be close to the worst-case scenario in real field top soil: (1) according to the documentation for plant protection products containing TBZ or CLP, the real application doses are 100e350 g ha1 for TBZ and 300e2600 g ha1 for CLP (CISTA, 2016); (2) considering 5 cm topsoil (relevant for epigeic earthworms) and a soil density of 1.5 g cm3 (FOCUS, 1997), the maximum doses correspond to 0.5 and 3.5 mg kg1 for TBZ and CLP, respectively; (3) the soil concentration might be higher at field hotspots, e.g. places where the application vehicle turns. Soils were contaminated according to the procedure of Doick et al. (2003). Briefly, the pesticides in acetone were added to the moist soils (at 50% WHC) in stainless steel bowls and mixed properly. After solvent evaporation in the fume hood, the water loss was measured by weighting of control (non-spiked) soils and

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Table 1 Properties of chlorpyrifos (CLP) and tebuconazole (TBZ) according to the Pesticide Properties DataBase (PPDB, 2016). Mw e molecular weight, Sw e solubility in water at 20  C, Kow e octanol-water partition coefficient, pKa e dissociation constant at 25  C, pv e vapour pressure at 25  C, Kh e Henry's law constant at 25  C, Koc e organic carbon-water partitioning coefficient, DT50 soil e typical soil degradation half-life under aerobic conditions. Structure

Pesticide type


Sw 1

log Kow




Koc 3


DT50 soil


Pa.m .mol






















Table 2 Physico-chemical properties of soil samples used in the study. TOC e total organic carbon content; HA e humic acids; FA e fulvic acids; pHKCl e exchangeable pH; particle size analysis: clay (<1 mm), fine silt (1e10 mm), silt (10e50 mm), fine sand (50e250 mm), and sand (0.25e2 mm), WHCmax e maximal water holding capacity. Properties



TOC (%) HA/FA pHKCl Clay (%) Fine silt (%) Silt (%) Fine sand (%) Sand (%) WHCmax

1.02 0.72 6.14 10.7 19.3 41.7 24.7 3.6 0.52

1.93 0.58 5.88 20.7 29.5 32.2 15.3 2.3 0.57

replenished. Contaminated soils were stored in airtight glass vessels in the dark at 22 ± 1  C for 4 days (recommended by OECD) until the beginning of uptake experiment (OECD, 2010), when both soils were mixed once again and a homogeneity of distribution of test substances was verified. The same day, earthworms were added and the experiment started (day 0). 2.3. Experimental design of bioaccumulation test A bioaccumulation test with E. andrei was performed to study the uptake kinetics of test compounds from soil. The experimental design was based on OECD guideline 317 (OECD, 2010). The soil amount (150 g of dry weight soil) was used for exposure to fill up 1/ 3 of the vessel's volume to allow aeration of soil. Each group of 8 earthworms from acclimatization was cleaned in tap water to remove soil particles, gently dried, weighted, and placed into the test vessel. The vessels were covered with a perforated lid. The experiment was carried at 22 ± 1  C in daylight, without feeding. Soil moisture was checked and readjusted by weighting the vessels twice a week. Sampling was performed after 0.5, 1, 3, 5, 7, 10, 14, 18 and 21 days of exposure. Three independent replicates (with 8 earthworms in each replicate) were used for each sampling point. Two replicates of solvent control (without test substances) were also prepared and sampled on the last day. Therefore 29 vessels were started for each soil upon the initiation of the test. After the exposure, earthworms from independent replicates were removed, cleaned in tap water, dried, weighed and placed in a Petri dish with moistened filter paper for 24 h to defecate. After depuration, earthworms were cleaned, dried, weighed, and frozen in 80  C until the analysis. Eight earthworms from the replicate were processed and analysed together as a cluster. The changes in the concentrations of substances in the soils


GUS leaching potential index 1






were monitored in parallel with the uptake kinetics of earthworms by measuring total concentrations of chemicals in soils at the same sampling points, including day 0. Soil was randomly taken from each vessel to get 5 gdw and frozen at 80  C until the analysis.

2.4. Determination of chemical concentrations in earthworms and soils QuEChERS extraction was used as the method to assess the total concentration of pesticides in soil and earthworm samples. This method has been repeatedly shown to be a fast, simple and effective approach to determine a broad spectrum of pesticides in environmental samples (Anastassiades et al., 2003; Lesueur et al., 2008; de Oliveira Arias et al., 2014; Yu et al., 2016). Frozen earthworms were lyophilized using a Christ Gamma 1e16 LSC Freeze Dryer (SciQuip, UK). Lyophilized weighed earthworms of one replicate (i.e. vessel) were put in polypropylene testtubes (50 ml) and homogenized by shaking with a ceramic homogenizer. A recovery standard of metolachlor (Sigma-Aldrich, Germany) was added before extraction (100 ml, 20 mg ml1). Homogenized earthworm samples were moistened by distilled water (9.5 ml) and after addition of acetonitrile (9.9 ml) shaken for 1 min. Test-tubes were immersed for 15 min in an ultrasonic bath. Partition was encouraged by adding extraction kits (Agilent Technologies, US) with buffered extraction salts (4 g Magnesium Sulphate, 1 g Sodium Chloride, 1 g Sodium Citrate, 0.5 g Sodium Hydrogencitrate Sesquihydrate). Samples were again shaken for 1 min and centrifuged. After centrifugation, the acetonitrile layer containing the analysed compounds was collected and cleaned by 1min shaking in a dispersive solid phase extraction kit (Agilent Technologies, US) to remove lipids from samples (150 mg primary secondary amine and 900 mg Magnesium Sulphate in a 15 ml polypropylene test-tube). Elutes were analysed by LC-MS/MS. As internal standard of D-metolachlor was injected onto a chromatographic column together with each sample. The limits of detection of the method for CLP, TBZ and metolachlor were 0.33 ng ml1 and the limits of quantification were 1 ng ml1. The recoveries of metolachlor added before the extraction were used to adjust chlorpyrifos and tebuconazole concentrations. The average recovery of metolachlor in earthworm samples (n ¼ 52) was 101 ± 5%. Coefficients of variation of earthworm samples (n ¼ 3) ranged from 2 to 28% for CLP and from 1 to 28% for TBZ. The total concentrations of test compounds in the soil samples were analysed using a similar procedure as for the earthworm samples but without the steps of homogenization and lipid removal. The average recovery of metolachlor in the soil samples (n ¼ 66) was 97 ± 4%. Coefficients of variation of soil samples

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(n ¼ 3) ranged from 6 to 34% for CLP and from 2 to 16% for TBZ. 2.5. Kinetic modelling and statistical analysis Kinetic modelling: The following first-order kinetic equation was used to describe a decrease in the concentration of tested compounds in soil during uptake (OECD, 2010):

  Cs ¼ C0 ek0 t


where Cs is the concentration of the substance in the soil (ng.gsoil1 dw ), C0 is the initial concentration of the substance in soil (ng.gsoil1 1 dw ), k0 is the degradation rate constant in soil (d ), and t is the time (in days). The changes in soil concentration were also considered in modelling the uptake kinetics. A one-compartment model was used to describe the uptake kinetics of pesticides by earthworms and to estimate the uptake and elimination rate constants (OECD, 2010):

Ca ¼

 ks     C0  k0 t  e   eke t ke  k0


where Ca is the concentration of the substance in worms 1 (ng.gearthworm-dw ), ks is the uptake rate constant in tissue (gsoil1 1 1 dw.gearthworm-dw.d ), and ke is the elimination rate constant (d ). The models were fitted to the raw data (time, concentrations) and degradation, uptake, and elimination rate constants (k0, ks, and ke) were calculated using the nonlinear estimation module of STATISTICA 13 (Dell Inc, 2015) with the LevenbergeMarquardt estimation method for the calculation of least squares. Calculation: The bioaccumulation factors (BAFss) were calculated as mean ± standard error (SE) obtained by dividing the concen1 tration in the earthworms Ca (ng.gearthworm-dw ) in each replicate (vessel) by the concentration in the soil Cs (ng.g1 soil-dw) in the same replicate (vessel) at the steady state (three time-points not significantly different) defined according to OECD guideline 317 (OECD, 2010):

BAFss ¼  

Ca Cs


The dynamic bioaccumulation factor (BAFd) was calculated by dividing the uptake rate constant in tissue, ks (gsoil-dw.gearthworm-dw-1.d-1) and the elimination rate constant, ke (d-1):

BAFd ¼  

ks ke


The SE of BAFss and BAFd were calculated from the SE of Ca, Cs, ks and ke using the propagation of error rule (Harvard University, 2017):

sffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi     SEA 2 SEB 2 SEC   ¼  C      þ  A B


where C stands for BAFss or BAFd, A stands for Ca or ks, respectively, and B stands for Cs and ke, respectively. The bioconcentration factor (BCFm) was modeled according to the EFSA Guidance document “Risk assessment for birds and mammals” (EFSA, 2009) and originates from the work of Jager (1998):


0:84  þ  0:012 Kow ¼ f oc     Koc


where Kow is the octanol/water partition coefficient, foc is the

organic carbon content of soil and Koc is the organic carbon/water partition coefficient. The BCFm for earthworm was defined as the concentration in earthworms on their fresh weight to the concentration in soil on dry weight. Therefore, we calculated the dry/fresh weight ratio of the earthworms collected at the end of the exposure and divided our BCFm results by its mean value 0.2 to enable a comparison with our BAFs. Degradation half-life times (T1/2, days) were calculated from k0 (d1) with the equation:

T1=2 ¼ ln2=k0


Statistical analyses: All statistical analyses were performed in STATISTICA 13 (Dell Inc, 2015). The effect of time on total concentrations in soils and earthworms was tested by analysis of variance (ANOVA) with the Tukey HSD multiple comparisons test after homogeneity of variances between groups (timepoints) was confirmed by Levene's test (p > 0.05). If the variances were heterogeneous (Levene's test with p < 0.05), non-parametric KruskalWallis ANOVA was used to verify the results of the parametric ANOVA. Then, multiple comparisons between groups (timepoints) were performed by a nonparametric Mann-Whitney U test. Steady state (equilibrium), defined according to OECD guideline 317 (OECD, 2010), was also determined by ANOVA and multiple comparisons e the concentration of pesticide in earthworms was considered to reach steady state when the concentrations in earthworms in three successive samplings did not significantly differ (p > 0.05). Comparisons of rate constants (k0, ks, and ke) and BAF values between compounds and soils were done with T-test. For kinetics constants, errors of estimation revealed by the regression were used as their SE. 3. Results 3.1. Total concentration of compounds in soil The following initial CUP concentrations were measured on day 0 of the bioaccumulation experiment (mean ± standard deviation, n ¼ 3): 3.86 ± 0.55 and 6.01 ± 0.56 mg.kg1 soil-dw in FS1 and FS2, and CLP respectively, 4.74 ± 0.25 and 5.32 ± 0.72 mg.kg1 soil-dw in FS1 and FS2, and TBZ respectively. The model Eq. (1) used for the compound decrease in soils provided a significant but relatively poor model for the dissipation data (R ¼ 0.45e0.61, p < 0.05). This is due to the fact that both pesticides did not greatly decrease in the soil and there was also an apparent variability between triplicate subsamples taken from the bottle showing soil heterogeneity. R values of each model and degradation rate constants in soils (k0) are summarized in Table 3. Fig. 1 shows the curves representing the temporal decrease of total concentrations in soils during the experiment. The results were expressed in a relative scale to compare the degradation kinetics of compounds in both soils: concentrations in soils in each sampling

Table 3 Degradation parameters of chlorpyrifos (CLP) and tebuconazole (TBZ) in soils: degradation rate constants in soil during bioaccumulation test (k0; d1) and degradation half-lives (T1/2; days). R values indicate fit of the models. The model output values are shown with errors of the estimation (±SE). The values in columns followed by the same letter are not significantly different from each other (T-test with a ¼ 0.05).



k0 (d1)


0.0142 ± 0.0047 0.0232 ± 0.0039 0.0116 ± 0.0011 0.0053 ± 0.0018

(b) (a) (b) (c)


T1/2 (d)

0.47 0.50 0.61 0.45

48.8 29.3 59.8 130.8

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point were calculated as a percentage of the initial concentration (day 0). A larger decline in concentration over 21 days was observed for chlorpyrifos (20.4% in FS1, 27.3% in FS2) than for tebuconazole (15.9% in FS1, 10.2% in FS2). The significantly (T-test, p < 0.05) highest value of k0 was found for chlorpyrifos in FS2 (0.0232 ± 0.0039 d1) and the significantly (T-test, p < 0.05) lowest value of k0 was found for tebuconazole (0.0053 ± 0.0018 d1) in the same soil. The degradation rate constants of chlorpyrifos and tebuconazole in FS1 were not significantly different (T-test, p > 0.05). The degradation rate constant of CLP was significantly (T-test, p < 0.05) higher in soil FS2 than FS1, but on the contrary, a significantly (T-test, p < 0.05) higher degradation rate constant of TBZ was reached in FS1 than FS2.

3.2. Uptake of compounds by earthworms The mortality of earthworms at the end of the experiment (15 ± 9%) was comparable to the solvent control (13 ± 10%) which was slightly above the validation limit of 10% as required by the OECD but was acceptable considering the low number of earthworms (8) on replicate. No abnormalities were found either on earthworms exposed to test compounds or in the solvent controls. The mean weight of earthworms after the exposure decreased (compared to the initial weight) but this decrease did not exceed 20%, which is the value required by OECD guideline 317. A steady state (equilibrium), defined according to OECD guideline 317 (concentration in earthworms in three successive sampling did not significantly differ, ANOVA, p > 0.05), was reached after 3 days of exposure of chlorpyrifos in both soils and after 3 days of exposure in soil FS2 and 5 days in soil FS1 for tebuconazole, respectively. Mean BAFss ± SE calculated at the steady state can be seen in Table 4. To be able to compare soils in graphs directly, the concentrations in earthworms (Ca) were normalized to the initial concentrations in   and Hofman, 2014) (Fig. 2). Uptake and elimisoils (C0) (Smídov a nation rate constants (ks and ke) are independent of this transformation. The uptake kinetic model Eq. (2) described well the concentration changes in worms for all compounds and soils (R ¼ 0.89e0.96, p < 0.05). R values of the model Eq. (2) and particular values of uptake and elimination rate constants may be found in Table 4. The uptake rate constant of both compounds reached their highest values in FS1 (4.88 ± 0.42 and 4.98 ± 0.53 1 gsoil-dw.gearthworm-dw .d1 for CLP and TBZ, respectively) and significantly (T-test, p < 0.05) differ from those in FS2. There was no significant (T-test, p > 0.05) difference in ks values between the two compounds in each soil. The elimination rate constant of chlorpyrifos was significantly (T-test, p < 0.05) higher in FS1 than FS2, but on the contrary, the ke value of tebuconazole was significantly


(T-test, p < 0.05) higher in FS2 than FS1. Also, dynamic BAF (BAFd) was counted and its mean values ± SE are summarized in Table 4. Modeled bioconcentration factors (BCFm) of both pesticides were higher in soil FS1 (36.2 and 38.9 for CLP and TBZ, respectively) than FS2 (19.1 and 20.5 for CLP and TBZ, respectively).

4. Discussion 4.1. Total concentration of compounds in soil The main dissipation pathways of chlorpyrifos in soil discussed in the literature are microbial degradation, volatilization and abiotic hydrolysis (Awasthi and Prakash, 1997; Baskaran et al., 1999; Chai et al., 2013; Racke et al., 1996). In our study, a major role of volatilization and abiotic hydrolysis is assumed, because the soils were sterilized. Even if CLP is only mildly volatile with a vapour pressure (pv) of 1.4 mPa (25  C) and Henry constant (Kh) of 4.8  101 Pa m3.mol1 (25  C) (PPDB, 2016), the study of Getzin (1981) showed that the volatilization of CLP is the major dissipation pathway under moist soil conditions. PPDB (2016) states a typical CLP soil degradation half-life of 50 days. However, the reported half-lives of CLP in non-sterile soils under laboratory conditions in the literature vary from 1.1 (Spurgeon et al., 2011) to 1576 days (Racke et al., 1994) with 3,5,6-trichloro-2-pyridinol as the major degradation product. This large range in half-life has been attributed to variation in factors such as pH, temperature, moisture content, organic carbon content, clay content and application rate (Awasthi and Prakash, 1997; Baskaran et al., 1999; Chai et al., 2013; Cink and Coats, 1993; Fang et al., 2009; Racke et al., 1996). Our degradation rate constants of CLP were 0.0142 and 0.0237 d1 and half-lives 48.8 and 29.3 days in FS1 and FS2, respectively and these values are comparable with the results of Racke et al. (1996) where k0 in sterilized soil ranged from 0.002 to 0.063 d1 and the half-lives were 11e341 days. In another study with sterilized soil, CLP degraded even more slowly (k0 of 0.0027e0.0040 d1 and half-lives of 173e256 days) (Chai et al., 2013). The slightly faster degradation of CLP in FS2 might possibly be caused by claycatalysed hydrolysis due to higher clay content (Getzin, 1981; Awasthi and Prakash, 1997). Tebuconazole was dissipated mainly by microbial transformations, rather than abiotic degradation, in previous studies (Li ~ oz-Leoz et al., 2011; Wang et al., 2012) with four et al., 2015; Mun identified degradants which derived from hydroxylation of the parent compound and/or chlorophenyl ring cleavage (Strickland et al., 2004). This agrees with our results where the dissipation of CLP in both soils was larger and faster than TBZ probably due to additional pathways of volatilization and hydrolysis which have no part in the mainly microbial degradation of TBZ. Volatilization to air

Fig. 1. Decrease in the concentration of pesticides in soils (FS1 e dotted line and white spots, FS2 e solid line and black spots) during the bioaccumulation test adjusted to the percentage of the initial concentration (day 0). The lines show a model of nonlinear regression according to Eq. (1) made with the least square method.

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Table 4 1 Bioaccumulation parameters of chlorpyrifos (CLP) and tebuconazole (TBZ) in soils: uptake rate constants (ks; gsoil-dw. gearthworm-dw . d1), elimination rate constants (ke; d1), BAFd calculated by dividing ks and ke, BAFss calculated at steady state by dividing Ca and Cs. R values indicate fit of the models. The model output values are shown with errors of the estimation (±SE). The values in columns followed by the same letter are not significantly different from each other (T-test with a ¼ 0.05).





4.88 ± 0.42 2.58 ± 0.46 4.98 ± 0.53 1.96 ± 0.33

ke (a) (b) (a) (b)

0.623 ± 0.058 0.392 ± 0.078 0.332 ± 0.041 0.866 ± 0.156

(a) (b) (b) (a)




0.96 0.89 0.95 0.90

7.84 ± 0.99 (a) 6.56 ± 1.77 (ab) 15.02 ± 2.45 (a) 2.26 ± 0.56 (b)

6.34 ± 1.30 (b) 4.51 ± 0.76 (b) 13.15 ± 1.01 (a) 2.17 ± 0.06 (c)

Fig. 2. Uptake kinetics of pesticides by earthworms during the 21-day exposure in soils (FS1 e dotted line and white spots, FS2 e solid line and black spots). Concentrations in earthworms (Ca) were standardized, i.e. divided by the mean initial concentration of the studied pesticides in soils (C0), to enable comparison of different soils the same plot. Curves show models derived from nonlinear regression according to Eq. (2).

(slowly exchanging via holes in the lids) in the bottles is improbable as TBZ has pv of 1.3  103 mPa (25  C) and Kh of 1.0  105 Pa m3.mol1 (25  C) e it is a practically non-volatile compound. The half-life of TBZ ranges in the literature from 9 to 263 days ~ oz-Leoz et al., 2011) and a with a k0 from 0.0008 to 0.0528 d1 (Mun typical soil degradation half-life of 63 days (PPDB, 2016). Although the soils in the literature studies were not sterilized, these values are in a similar range as our results (half-lives of 60 and 131 days, 0.0116 and 0.0053 d1). The variation in values of triazole fungicides half-lives was ascribed to temperature, moisture content, pH, organic carbon content, concentration in soil and enantioselectivity ~ oz-Leoz of degradation (Bromilow et al., 1999; Li et al., 2015; Mun et al., 2011; Strickland et al., 2004; Wang et al., 2012). The slower degradation of TBZ in FS2 was probably caused by sorption of TBZ     et al., on SOM and clay minerals (Cadkov a et al., 2012, Cadkov a 2013) and thus a lower availability of TBZ for degradation. 4.2. Uptake of compounds by earthworms 4.2.1. Chlorpyrifos e equilibrium and BAFs Determination of BAF at the steady state (after an equilibrium in concentrations was reached between earthworm, soil and soil pore water) is one of the approaches for assessing bioaccumulation of chemicals in earthworms. In the study of Lister et al. (2011), a steady state for CLP was reached after 3 days of exposure similarly to our study. Equilibrium was also reached in the study of Spurgeon et al. (2011) during a 7-day-long exposure, but the exact time point was not mentioned. BAF values for CLP in the literature ranged from 0.22 to 0.731 (Lister et al., 2011; Yu et al., 2006) and were counted as the ratio of the concentration in earthworms on their fresh weight and the concentration in soil on dry weight, which is not comparable with our results. Therefore, we converted our results by the dry/fresh weight ratio of the earthworms (0.2) to enable a comparison of the data using earthworm fresh weights. Then, our BAFss for CLP are slightly higher in both soils (1.26 for FS1, 0.90 for FS2) than in the studies of Yu et al. (2006) and Lister et al. (2011), although the range

of OM content in soils is similar [factor 1.7 used to calculate OM from our TOC (Yu et al., 2012)]. The discrepancy might for example originate from different soil and SOM structure and composition or assimilation efficiency of the earthworm species used (Burgess et al., 2003). As can be seen from Table 4, dynamic BAF values (BAFd) for CLP are comparable with BAF values calculated with concentrations in the steady state (BAFss). The results show that the modeled BCFm values cannot really be compared with the BAF values acquired experimentally. The BCFm values many times overestimated the bioaccumulation potential of CLP. BCFm are higher in soil FS1 than FS2. The BAF values show the same trend but the ratio between them is not the same. 4.2.2. Tebuconazole e equilibrium and BAFs The time needed to reach a steady state of TBZ in the study of Yu et al. (2012) differed between its enantiomers and two different concentrations: 3 and 7 days for (þ)-R-tebuconazole which was comparable with our results (3 and 5 days in different soils). Equilibrium was not reached for ()eS-tebuconazole because of a peak-shaped uptake (peak 0.5e1 days) followed by a decrease in concentration. The accumulations of the five other triazole fungicides reached a plateau in about 11 days in the study of Chen et al. (2014). The longer time to reach a steady state of TBZ in FS1 was probably caused by the approximately two-times-higher bioavailability (based on BAF values) of TBZ in this soil and thus a higher equilibrium concentration than in FS2. Yu et al. (2012) calculated BSAFs after 36 days of exposure to express the bioavailability of TBZ enantiomers. A steady state was not reached for ()eS-tebuconazole, but for (þ)-R-tebuconazole the equilibrium was achieved after 3e7 days; this is good for an interstudy comparison, but this approach neglects the possible reduction of the bioavailable fraction by the ageing process. Even though our BAFss for TBZ (2.6 for FS1, 0.44 for FS2) are in the same range as the BAFs converted from their BSAFs: (þ)-R-TBZ e 1.394, ()eS-TBZ e 0.425. Yu et al. (2012) also counted an uptake and elimination constants ratio ((þ)-R-TBZ e 1.402, ()eS-TBZ e 0.831) which is comparable with our dynamic BAFd for TBZ (3.00 for FS1,

 et al. / Environmental Pollution 236 (2018) 257e264 M. Svobodova

0.46 for FS2). Our BAFd values for TBZ are, just like in the case of CLP, comparable with the BAFss values (Table 4) which suggests that calculating BAF as the ratio of uptake and elimination constants could be a suitable approach for bioaccumulation assessment. The BCFm values of TBZ show the similar trends as the modeled BCFs of CLP. The differences between the modeled BCFs and real BAFs could be caused by factors influencing bioaccumulation of these pesticides other than those included in the model (e.g. other soil properties than OC, exposure route). 4.2.3. Influence of soil properties Bioaccumulation of organic pollutants in soils is strongly dependent on their bioavailability, which is a function of soil, the compound properties and also the organism's ecology and physiology. The major soil properties affecting bioavailability of CUPs are a and Papini, 2005; Papini et al., 2006; Wu SOM, clay and pH (Andre et al., 2011a; Yu et al., 2006). The adsorption capacity of soil for the pesticides chlorpyrifos, butachlor and myclobutanil (triazole fungicide) positively correlated with SOM content in the study Yu et al. (2006), while according to Felsot and Dahm (1979), adsorption of chlorpyrifos occurred mainly through hydrophobic in  teractions with organic matter surfaces. The results from (Cadkov a   et al., 2012, Cadkov a et al., 2013) showed the importance of SOM content and composition for tebuconazole sorption, but also the contribution of clay minerals and Fee and Mn-oxyhydroxides. Our results for CLP showed slightly lower bioavailability (according to BAFs) and slower uptake in soil with a higher TOC content (FS2) where the availability of CLP for uptake by earthworms by dermal passive diffusion from pore water might be decreased by sorption of CLP on SOM (soil FS2 had twice as much TOC as FS1). The same mechanism might cause the lower bioavailability and slower uptake of TBZ in FS2 than in FS1. On top of SOM driven sorption, sorption to clay minerals might contributed to the lower bioavailability of TBZ in FS2. TBZ sorption on clay minerals has been re  ported for TBZ (Cadkov a et al., 2013) but not for CLP. The higher elimination rate of TBZ in FS2 than in F1 might be a result of the higher TOC content in soil too. SOM content is considered to be an influential factor in the elimination of hydrophobic organic compounds by earthworms (Belfroid and Sijm, 1998). Faster elimination rates in soil with a higher OM content also occurs in the studies of de Barros Amorim et al. (2002) (lindane), Belfroid and Sijm (1998) (PeCB, HCB) and Paul and Ghosh (2011) (PCBs). We have no explanation for the faster elimination of CLP in FS1 than FS2. 4.2.4. Influence of compound properties Among CUP properties, Kow, solubility and acid dissociative constant (pKa) are the most important in influencing their soil behaviour and bioavailability (Chang et al., 2016; Felsot and Dahm, 1979; Xu et al., 2014; Yu et al., 2006). CLP has a low water solubility of 1.1 mg.L1 (25  C), log Koc 3.9, log Kow 4.7 (PPDB, 2016) and hence it sorbs to SOM. TBZ was more bioavailable than CLP in FS1, which corresponded with the higher water solubility of TBZ of 36.0 mg.L1 (25  C) and lower values of log Koc 2.9 and log Kow 3.7 (PPDB, 2016) which suggests lower sorption on SOM and greater availability via pore water. However, CLP was more bioavailable and its uptake faster in FS2 than TBZ. Sorption to soil particles makes contaminants less bioavailable via the pore-water exposure route but more available via the ingestion exposure route. The most likely explanation for the higher and faster uptake of CLP is that the major uptake route of CLP was via active ingestion of pesticide together with soil (Yu et al., 2006). Belfroid et al. (1995) concluded that food intake becomes a relevant additional uptake route for very hydrophobic chemicals (log Kow > 5). The higher log Kow value of 4.7 for


CLP may result in higher effective accumulation in earthworms after being ingested. Another explanation is binding of TBZ on clay    et al., 2012, Cadkov  et al., 2013), decreasing its minerals (Cadkov a a availability in FS2 which had a higher content of clay. 5. Conclusions According to our results, an equilibrium between concentrations of pesticides in earthworms and soil was reached after 3 days of exposure for chlorpyrifos in both soils and for tebuconazole in soil FS2. In FS1 a five-day exposure was needed to establish the equilibrium for tebuconazole which was probably caused by an approximately two-times-higher bioavailability of TBZ in FS1 and thus a longer time required to achieve equilibrium concentration. In our study, BAFs calculated in steady state were comparable with dynamic BAFs. Therefore, we presume that determination of BAF as a ratio of uptake and elimination rate constants is a reliable approach for assessing bioavailability. The modeled BCFs overestimated bioaccumulation of both pesticides in earthworms probably due to other factors influencing bioaccumulation and availability for uptake of these pesticides that were not included in the model. The results suggested that degradation of CLP and TBZ during exposure was influenced by SOM (TOC) and clay content but in a different way. In both cases higher SOM content probably reduced the availability of CLP and TBZ for degradation. However, the higher clay content possibly contributed to degradation of CLP due to clay-catalysed hydrolysis. On the other hand, higher clay might bring more abundant sorption sites for TBZ to reduce its availability for degradation. Soil properties also affected bioaccumulation of tested pesticides by earthworms. Availability of CLP and TBZ for uptake by earthworms was likely decreased by sorption on SOM and TBZ binding on clay minerals. The results also imply that the properties of the tested pesticides play a role in their bioavailability and bioaccumulation. In soil FS1, the higher water solubility and lower Koc and hydrophobicity of TBZ compared to CLP led to higher availability of TBZ through pore water exposure. In soil FS2, the higher hydrophobicity of CLP probably caused its additional uptake via ingestion and thus higher bioavailability of CLP than TBZ for earthworms. Acknowledgment The current study was funded by the Czech Science Foundation  (GACR) grant number 15-20065S, which is gratefully acknowledged. This work was also partly supported by the Czech Ministry of Education, Youth and Sports of the Czech Republic (project M2015051 and project CZ.02.1.01/0.0/0.0/16_013/0001761 RECETOX RI). References  Anastassiades, M., Lehotay, S.J., Stajnbaher, D., Schenck, F.J., 2003. Fast and easy multiresidue method employing acetonitrile extraction/partitioning and “dispersive solid-phase extraction” for the determination of pesticide residues in produce. J. AOAC Int. 86, 412e431. a, M.M.D., Papini, S., 2005. Influence of soil properties on bioaccumulation of Andre 14C-simazine in earthworms Eisenia foetida. J. Environ. Sci. Heal B 40, 55e58. Awasthi, M.D., Prakash, N.B., 1997. Persistence of chlorpyrifos in soils under different moisture regimes. Pest Manag. Sci. 50, 1e4. de Barros Amorim, M.J., Sousa, J.P., Nogueira, A.J., Soares, A.M., 2002. Bioaccumulation and elimination of 14C-lindane by Enchytraeus albidus in artificial (OECD) and a natural soil. Chemosphere 49, 323e329. Baskaran, S., Kookana, R.S., Naidu, R., 1999. Degradation of bifenthrin, chlorpyrifos and imidacloprid in soil and bedding materials at termiticidal application rates. Pest Manag. Sci. 55, 1222e1228. Belden, J.B., Phillips, T.A., Coats, J.R., 2004. Effect of prairie grass on the dissipation, movement, and bioavailability of selected herbicides in prepared soil columns. Environ. Toxicol. Chem. 23, 125e132. Belfroid, A., Meiling, J., Drenth, H.J., Hermens, J., Seinen, W., van Gestel, C.A.M., 1995.


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