Bioaccumulation of atrazine and chlorpyrifos to Lumbriculus variegatus from lake sediments

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Ecotoxicology and Environmental Safety 71 (2008) 860–868 www.elsevier.com/locate/ecoenv

Bioaccumulation of atrazine and chlorpyrifos to Lumbriculus variegatus from lake sediments$ A.P.K. Jantunen, A. Tuikka, J. Akkanen, J.V.K. Kukkonen Faculty of Biosciences, Laboratory of Aquatic Ecology and Ecotoxicology, University of Joensuu, Yliopistokatu 7, P.O. Box 111, 80101 Joensuu, Finland Received 5 September 2007; received in revised form 10 January 2008; accepted 19 January 2008 Available online 18 March 2008

Abstract The bioaccumulation of the pesticides chlorpyrifos and atrazine to the benthic oligochaeta Lumbriculus variegatus from four diverse artificially contaminated lake sediments (OC 0.13–21.5%) was studied in the laboratory. The steady state of bioaccumulation was not reached within 10 d. Chlorpyrifos showed stronger bioaccumulation than the less lipophilic atrazine, the biota-sediment accumulation factors (BSAFs) being 6.2–99 for the former and 1.9–5.3 for the latter. While bioaccumulation factors (BAFs) dropped with increasing organic content of the sediments, the high level and considerable range of the obtained BSAFs indicate other sediment qualities, such as the age and characteristics of the organic material, having a strong effect on the bioavailability of these compounds. The slow and incomplete desorption of chlorpyrifos from the most inorganic sediment indicates also that this compound may be strongly bound to some type of inorganic material. Any specific influential sediment fraction or characteristic could not be identified. r 2008 Elsevier Inc. All rights reserved. Keywords: Chlorpyrifos; Atrazine; Sediment; Lumbriculus variegatus; Bioaccumulation; Desorption

1. Introduction Sediments are a known sink for hydrophobic chemicals. Through leaching and runoff, water systems tend to accumulate chemicals from the entire watershed, and any hydrophobic chemical tends to attach to organic or particulate matter (including living organisms) and to eventually settles with it to the bottom. In sediment, hydrophobic chemicals may become associated with either organic or inorganic material or dissolved in interstitial water. Absorption, which is considered a linear and fully reversible process, occurs to young organic material consisting of partly degraded or reconstituted biopolymers, lipoproteins, amino acids, lipids, and humic or fulvic substances (Cornelissen et al., 2005a). Adsorption, apparently a nonlinear and competitive process, can occur to carbonaceous geosorbents (‘‘black carbon’’, BC) (Cornelissen et al., 2005a) as well as some inorganic materials such as $

No human subjects or vertebrate animals were involved in this study.

Corresponding author. Fax: +358 13 251 3590.

E-mail address: [email protected].fi (A.P.K. Jantunen). 0147-6513/$ - see front matter r 2008 Elsevier Inc. All rights reserved. doi:10.1016/j.ecoenv.2008.01.025

clays. While absorption seems to lead to reasonably predictable concentration-dependent equilibria between the sorbed and the dissolved state, adsorption can lead to extensive and largely irreversible sorption at low aquatic concentrations when the maximum adsorption capacity of the sorbent has not been reached. Benthic organisms may become contaminated by chemicals in sediments, either by absorbing them from ingested sediment in the gastrointestinal tract or across the body wall from pore water or water overlying the sediment. The relatively easy desorption of hydrophobic chemicals absorbed in the equilibrium-partitioning (EqP) domain of the sediment seems to mean predictable bioavailability of the chemical to benthic organisms, with a fairly constant coefficient of partitioning to the organic content of the sediment (KOC) and therefore also a fairly constant biota-sediment accumulation factor (BSAF) (Di Toro et al., 1991). On the other hand, the bioavailability of hydrophobic chemicals when a nonlinear sorption domain is present can be very limited (Cornelissen et al., 2005a), and is currently difficult to predict on basis of the characteristics of the sediment and the chemical.

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However, a correlation has been found between the bioavailability (BSAF) and desorption parameters (generally the rapidly desorbed fraction of the chemical, Fr, according to the two- or three-compartment first-order desorption model (Cornelissen et al., 1997b)) of polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs) and chlorobenzenes in sediment (Kraaij et al., 2001; Leppa¨nen et al., 2003; ten Hulscher et al., 2003; Kukkonen et al., 2004). The rapid desorption (Kukkonen et al., 2003) and bioavailability (Kukkonen et al., 2005) of planar compounds have been found to correlate positively with plant-derived ‘‘young’’ carbon in the sediment (lignin, pigments), while those of nonplanar compounds seem to simply correlate negatively with organic carbon, indicating equilibrium partitioning. The triazine herbicide atrazine and the organophosphorus insecticide chlorpyrifos are used globally in large quantities in crop protection. Both pesticides may contaminate water systems through spray drift or runoff from agricultural fields (US Environmental Protection Agency, 2002, 2003) and may persist in sediments (Brock et al., 1992; US Environmental Protection Agency, 2003; Bondarenko and Gan, 2004). Chlorpyrifos is considered moderately to very highly toxic to fish and aquatic invertebrates (US Environmental Protection Agency, 2002), and atrazine slightly to moderately toxic to freshwater fish and slightly to highly toxic to freshwater invertebrates as well as persistent in aquatic environments (US Environmental Protection Agency, 2003). Both compounds have a nonplanar molecular structure, although the lowest-energy structure of the atrazine molecule is nearly planar (Cornelissen et al., 2005b). In order to add to the little current knowledge about the consequences of the chronic exposure of benthic organisms to sediment-associated atrazine or chlorpyrifos, we studied the bioaccumulation of atrazine and chlorpyrifos from four artificially contaminated sediments with a wide range of organic contents to Lumbriculus variegatus, a benthic oligochaete species commonly used in freshwater toxicity testing. 2. Materials and methods 2.1. Test organisms The L. variegatus Mu¨ller (Oligochaeta) worms used in the experiments originated from a laboratory culture maintained at the Laboratory of Aquatic Ecology and Ecotoxicology, University of Joensuu as described by Ma¨enpa¨a¨ et al. (2003).

2.2. Chemicals 14 C-labelled atrazine (2-chloro-4-ethylamino-6-isopropylamino-1,3,5triazine-ring-ul-14C; Sigma-Aldrich, St. Louis, Missouri, USA) and chlorpyrifos (pyridine-2,6-14C; American Radiolabeled Chemicals Inc., St. Louis, Missouri, USA) were used. Chlorpyrifos was dissolved in acetone (Merck, pro analysi) and atrazine in 96% ethanol. The basic properties of these pesticides are given in Table 1.

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Table 1 The physico-chemical characteristics of atrazine and chlorpyrifos (Mackay et al., 2006)

Molecular weight Molecular formula Log Kow Solubility in water (mg L1) pka Vapor pressure (Pa)

Atrazine

Chlorpyrifos

215.68 C8H14ClN5 2.00–2.80 28.0–30.0 (20 1C) 1.62–1.70 4  105 (20 1C)

350.59 C9H11Cl3NO3PS 4.70–5.27 0.73 (20 1C) N/A 52150  105 (20 1C)

Other chemicals used were the LSC cocktails Insta-Gels Plus and TM Ultima Gold from Packard Bioscience (Groningen, The Netherlands), s Soluene 350 tissue solubilizer (Packard Instrument BV, Groningen, The Netherlands), analytical-grade acetone from Merck (Darmstadt, Germany), PESTANAL-grade n-hexane from Sigma-Aldrich Laborchemikalien (Seelze, Germany), analytical reagent grade cyclohexane from Labscan (Dublin, Ireland) and HgCl2 from J.T. Baker (Deventer, The Netherlands).

2.3. Exposure media Four sediments (S1–S4) with differing properties (Table 2), originating from three lakes in Eastern Finland, were used. The sampling of these sediments has been described by Cornelissen et al. (2004). The total organic content (TOC) of the sediments was determined with an elemental analyzer (Model 1106, Carlo Erba Strumentazione, Milano, Italy), with which the nitrogen and hydrogen contents of the sediments were also analyzed. The humic organic content (OChumic) was determined from NaOH extracts with a TOC analyzer (Shimadzu TOC-5000A, Shimadzu Scientific Instruments Inc., Kyoto, Japan). The protein and pigment contents were determined spectrophotometrically (Hitachi U-2000 Spectrometer, Hitachi High-Technologies, Tokyo, Japan) from NaOH and ethanol extracts of the sediment, respectively (SFS (Finnish Standards Association), 1993 and Rausch, 1981, respectively). The lipid content was determined gravimetrically from trichloroethanol extracts by a modification of a method for lipid extraction from tissue samples by Parrish (1999). The black-carbon (BC) contents were determined by Cornelissen et al. (2004). Before use in these experiments, the sediments were sifted through a 1-mm sieve and homogenized by stirring. Sediments from the same locations have been used previously in sorption, bioaccumulation, and toxicity studies (Ma¨enpa¨a¨ et al., 2003; Jantunen and Kukkonen, 2003; Cornelissen et al., 2004; Ma¨enpa¨a¨ and Kukkonen, 2006). The remote location of the lakes and some earlier analyses suggest that the contaminant levels of the sediments are very low (Ristola et al., 1996, 1999). Aerated artificial fresh water was used throughout the experiments. The artificial freshwater was deionized water containing 1.0 mM inorganic salts (0.80 mM CaCl2  2H2O, 0.20 mM MgSO4  7H2O, 0.03 mM KCl and 0.31 mM NaHCO3). Its pH was adjusted to 7.0 using HCl and NaOH.

2.4. Bioaccumulation experiments The bioaccumulation of atrazine and chlorpyrifos was studied in the four different sediments, spiking a single compound to a single sediment at a time. Spiking was performed with 2.5 kg of sediments in a 4-L beaker, adding the [14C]-labelled compound to give a final concentration of ca. 5 nmol/g of organic carbon in S1–S3 and ca. 30 nmol/g (OC) in S4, which has a very low organic content. The spiked sediments were mechanically mixed with a rotating metal blade for 3 h at room temperature. The contact time before the start of experiments was 5 d, of which 2 d was at +7 1C and 3 d at room temperature for the bioaccumulation experiment. The bioaccumulation experiments were performed at room temperature (2071.5 1C) with 16/8 h light/dark periods. A cold fluorescent light source

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Table 2 The characteristics of the test sediments S1–S4 (average7SD). LOI ¼ loss of ignition

Dry weight % LOI % TOC % OChumic % N% C:N H:C Proteins (mg g1 dw) Lipids (mg g1 dw) Pigments (mg g1 dw) BC:TOC %a a

S1

S2

S3

S4

11.01 36.7 21.54 23.17 1.05 20.5 0.12 92.35 (10.85) 10.822 (0.089) 94.18 (14.38) 0.4

14.27 9.3 3.47 3.64 0.21 16.5 0.23 1.99 (0.30) 0.518 (0.013) 16.92 (0.98) 3.2

42.83 3.1 1.65 1.00 0.05 33 0.17 3.36 (0.19) 0.731 (0.003) 19.53 (0.53) 11

75.06 0.4 0.13 0.17 0.00 – 0.31 1.19 (0.11) 0.283 (0.192) 3.95 (0.24) 20

Cornelissen et al. (2004).

(o500 nm) was used in order to prevent photodegradation. Evaporated water was replaced with deionized water when appropriate. The experimental units were 250-mL glass beakers containing ca. 60 g (wet weight) spiked sediment and 150 mL artificial fresh water. Water was carefully poured on top of the sediment, and the units were left to stabilize in experimental conditions for 3 d before the experiments were started. At the start of the experiment, eight L. variegatus were added into each unit. At each sampling time (6, 24, 48, 96, 144, and 240 h), three random replicate units were destructively sampled. Two 8-mL water samples were taken from the water column of each unit into 20-mL plastic LSC vials, 8 mL LSC cocktail (InstaGel) was added to the samples and the vials were shaken vigorously in order to form a gel. The rest of the water column was then carefully poured out of the beaker, and one sediment sample for dry weight (1–2 g wet weight) and two LSC samples (ca. 0.1 g wet weight) were taken. Before taking the sediment samples, the sediment was carefully stirred with a spoon except in the case of S4, since stirring this extremely sandy sediment damaged the animals. The dry-weight sediment samples were placed in pre-weighed aluminum foil cups, weighed for wet weight, dried in an oven (Memmert UM 500, Memmert GmbH, Schwabach, Germany) at 105 1C for 24 h, cooled to room temperature and weighed for dry weight. The LSC sediment samples were placed in 20-mL LSC vials and weighed, after which 0.5 mL tissue solubilizer was added and the samples were incubated overnight at 50 1C in an oven (Memmert UM 200, Memmert GmbH, Schwabach, Germany). Twelve milliliters of LSC cocktail (UltimaGold) was then added and the sample thoroughly mixed with Vortex. The animals from each unit were collected from the sediment by sieving, counted, and kept in clean water for 6 h before weighing and analysis in order to purge their gastrointestinal tract. They were then carefully blotted dry on tissue paper, placed on a pre-weighed piece of aluminum foil and weighed (Sartorius 4503 Micro, Go¨ttingen, Germany), and then placed in a 6-mL LSC vial with 0.5 mL tissue solubilizer, in which they were left at room temperature for at least 24 h before adding 6 mL LSC cocktail (UltimaGold) and thorough mixing with Vortex. Winspectral 1414 Liquid Scintillation Counter (LSC) from Wallac (Turku, Finland) was used for analyzing the scintillation of the samples. All samples were stored at room temperature for at least 2 d before analysis. The LSC data were corrected for background and quench by external standards.

weight) each, were extracted for the parent chemical. The samples were extracted twice with 5 mL cyclohexane, sonicated for 2 min (Vibracell, Sonics & Materials Inc., Danbury, USA) and also mixed thoroughly with Vortex after the first extraction, then centrifuged (Heraeus Multifuge 1 S-R, Thermo Electron Corporation, Osterode, Germany) at 3000 rpm for 5 min and the overlying solvent carefully pipeted into a LSC vial. The solvent extracts were combined and evaporated to a volume of 2 mL, 10 mL LSC cocktail (UltimaGold) was added and the samples were mixed thoroughly with Vortex, then stored for at least 48 h before analysis with LSC.

2.6. Lipid content The lipid content of the animals in each of the studied sediments was monitored during the bioaccumulation study with separate experimental units. Twenty-five grams of clean sediment was placed in 50-mL beakers, which were then filled with artificial fresh water (ca. 25 mL) and left to stabilize for 3 d in the experimental conditions. At the start of the experiment, four L. variegatus were added into each beaker and four times four animals were sampled directly from the laboratory culture in order to establish the starting level. At 120 and 240 h, four random units were sampled and the animals collected by sieving, counted, and weighed as in the bioaccumulation study. The animals from each unit were then frozen in a sealed glass tube for later lipid extraction. The lipid extraction method used was based on a total lipid determination method by Parrish (1999). The tissue samples were extracted three times with a 8:4:3 chloroform–methanol–water mixture, including the water content of the samples (for L. variegatus, 83.3% of the fresh weight) in the water fraction. The chloroform fractions (3  266.6 mL) were combined and evaporated from a pre-weighed aluminum foil cup, which left the lipids extracted from the tissue sample to be weighed. The temperature and oxygen contents of the water column were also monitored from these beakers (chlorpyrifos: 2, 96, 144, and 240 h, atrazine: 24, 120, and 240 h) using Multi-Line P4 Universal Meter (WTW, Weilheim, Germany).

2.7. Egestion rate 2.5. Degradation Separate experimental units parallel to those used in the bioaccumulation study were kept and destructively sampled during the bioaccumulation study in order to evaluate the degradation and biotransformation of the compounds. Sediment samples from these units were taken and stored in a freezer (20 1C) until extraction. With both chemicals and each sediment, one sediment sample taken at the beginning of the study and samples from duplicate units at the end of the study (240 h), 1 g (wet

The egestion rate of the animals in each of the studied sediments was monitored during the bioaccumulation study with separate experimental units. Twenty-five grams of spiked sediment was placed in 50-mL beakers, which were then filled with water (ca. 25 mL) and left to stabilize for 3 d in experimental conditions; however, after the beakers had settled for 2 d, a 2–3-mm layer of combusted quartz sand was carefully added on the sediment surface. At the start of the experiment, four L. variegatus were added into each beaker. At sampling times (approximately 24, 48, 72, 120, 158, 216, and 240 h), the fecal pellets left by the animals on the surface of

ARTICLE IN PRESS A.P.K. Jantunen et al. / Ecotoxicology and Environmental Safety 71 (2008) 860–868 the quartz sand were carefully collected with a glass pipette. The pellets were then filtered on fiberglass filters, dried to a constant weight at 105 1C, and weighed for dry weight.

2.8. Desorption experiments Desorption kinetics were studied with chlorpyrifos only, as a pilot experiment had indicated that the method was not suitable for atrazine. All four sediments were studied in duplicate. The experiment was started simultaneously with the bioaccumulation experiment, using the same spiked sediments after a 5 d contact time at +7 1C. In a slightly modified version of the Tenax method (Cornelissen et al., 1997a), 2–3 g spiked sediment, ca. 50 mL artificial fresh water with a biocidal 50 mg L1 concentration of HgCl2, and ca. 0.2 g Tenax beads (60–80 mesh TenaxsTA) were placed in 50-mL glass tubes and continuously mixed with a mechanical shaker (LD-76, Labingo BV, The Netherlands). The Tenax beads were collected for analysis and replaced at 3, 9, 24, 36, 48, and 72 h and 5, 7, 11, 16, and 24 d. The collected Tenax beads were extracted with 5 mL acetone by incubating them overnight at 50 1C (Memmert UM 200), after which the overlying acetone was carefully pipeted into a LSC vial. The Tenax beads were then extracted twice with 5 mL n-hexane for 1 h. The extraction solvents were combined and evaporated to a volume of 1–3 mL. Ten milliliters of LSC cocktail (InstaGel) was then added, and all samples were thoroughly mixed with Vortex and stored for at least 2 d at laboratory temperature before analysis with LSC.

2.9. Data analysis and kinetic models

3. Results 3.1. Bioaccumulation experiments No mortality of L. variegatus was observed during the experiments. Water temperature remained within 207 1.5 1C, and the measured oxygen content was 4.5– 6.8 mg L1 during the atrazine experiment and 5.5–7.5 mg L1 during the chlorpyrifos experiment. Decrease in sediment concentrations during the 10-day experiments was significant (and therefore taken into account when modelling the bioaccumulation data) in the case of atrazine in S1–S3 and chlorpyrifos in S2: atrazine concentrations dropped by 5%, 37%, and 13% in S1, S2, and S3, respectively, and the chlorpyrifos concentration by 14% in S2. Atrazine concentrations in the water column increased slightly through the experiment, while chlorpyrifos concentrations remained the same or decreased slightly. The bioaccumulation data and modelled bioaccumulation curves are presented in Fig. 1. The bioaccumulation of both atrazine and chlorpyrifos was relatively slow from all sediments, and the steady state was not reached for either compound within 240 h, except possibly in the case of chlorpyrifos in S4.

The bioaccumulation and desorption data were modelled with Micromaths Scientists software. The desorption data were fit with two- and three-compartment first-order desorption models, St ¼ F r expðkr  tÞ þ F s expðks  tÞ S0

863

S1

S2

S3

S4

0.30 Atrazine

(1)

0.25

and

where St and S0 are the chemical concentrations left in sediment at time t (h) and at the start of the experiment, Fr, Fs, and Fvs are the rapidly, slowly, and very slowly desorbed fractions of the chemical, and kr, ks and kvs the respective desorption rate constants. The fits of these models were compared with the extra sum-of-squares F-test (Motulsky and Christopoulos, 2003): F¼

ðSS1  SS2Þ=ðDF1  DF2Þ , SS2=DF2

(3)

where SS1 and SS2 are the sums of squares and DF1 and DF2 the degrees of freedom of the simpler (1) and the more complicated (2) model. The bioaccumulation data were fit with a first-order kinetic model (Landrum, 1989): Ca ¼

ks C ðt¼0Þ s ðelt  eke t Þ, ðke  lÞ

0.20

(2)

(4)

where Ca and Cs are the concentrations of the chemical in the animal and sediment (mmol kg1), ks is the uptake clearance coefficient (g (dw) g1 (ww) h1), ke is the conditional elimination rate coefficient (h1), and t is time (h). The decrease of sediment concentration was taken into account whenever the slope (l) of the observed sediment concentrations as a function of time differed significantly from zero. Parameters describing bioaccumulation (bioaccumulation factor, BAF, and biota-sediment accumulation factor, BSAF) were calculated in three different ways: BAF ¼ ks/ke, BAF ¼ C(ta ¼ 240h)/C(ts ¼ 0h) (in which Ca was expressed as mmol kg1 sediment (dw) and Cs as mmol kg1 tissue (ww)), and BSAF ¼ BAF normalized to the lipid content in the organism and the organic carbon content in the sediment.

0.15 Tissue concentration (µmol kg-1 ww)

St ¼ F r expðkr  tÞ þ F s expðks  tÞ þ F vs expðkvs  tÞ, S0

0.10 0.05 0.00 5 Chlorpyrifos 4

3

2

1

0 0

50

100

150 Time (h)

200

250

Fig. 1. Bioaccumulation of atrazine and chlorpyrifos in L. variegatus from sediments S1–S4.

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Table 3 Toxicokinetic parameters for L. variegatus exposed to atrazine and chlorpyrifos in sediments S1–S4 Sediment dose

Atrazine, S1 Atrazine, S2 Atrazine, S3 Atrazine, S4 Chlorpyrifos, Chlorpyrifos, Chlorpyrifos, Chlorpyrifos,

S1 S2 S3 S4

mmol kg1 (dw)

mmol kg1 OC

1.06 0.111 0.0580 0.0132 1.10 0.154 0.0724 0.0596

4.35 3.20 3.52 10.2 5.11 4.44 4.39 45.9

Lambda value

ks (g (dw) g1 (fw) h1)

ke (h1)

r2

0.0003 0.0018 0.0005 – – 0.0007 – –

0.00094 (0.00007) 0.00233 (0.00010) 0.01031 (0.00055) 0.11471 (0.02142) 0.0115 (0.0006) 0.0866 (0.0072) 0.1119 (0.0085) 0.7938 (0.1944)

0.00504 (0.00091) 0.00424 (0.00050) 0.00267 (0.00058) 0.00586 (0.00243) 0.0037 (0.0006) 0.0031 (0.0009) 0.0044 (0.0009) 0.0162 (0.0053)

0.991 0.996 0.995 0.947 0.996 0.988 0.991 0.911

The toxicokinetic parameters for L. variegatus exposure to atrazine and chlorpyrifos are presented in Table 3. The uptake clearance coefficients (ks) were the lowest for sediment S1 and increased through S2 and S3 to S4 for both compounds. The elimination rate coefficients (ke) were similarly highest in S4, but there is no clear trend through the other sediments. According to all bioaccumulation parameters (Table 4), chlorpyrifos has in each sediment a clearly higher bioaccumulation potential than atrazine. For both compounds, BAFs increased from sediment S1 through S2 and S3 to S4, while BSAFs had approximately the opposite trend. The two sets of BAFs calculated in different ways (Ca/Cs and ks/ke) seem to be in reasonable agreement, although the modelled BAFs are generally higher. 3.2. Degradation Of the chemical extractable from sediment at the beginning of the experiment (0 h), the extractable fractions at 240 h (assumed to represent the parent compound) were 55%, 94%, 51%, and 35% for atrazine and 99%, 85%, 95%, and 103% for chlorpyrifos from sediments S1, S2, S3, and S4, respectively. The total recoveries from different sediments were 67–108% for atrazine and 76–119% for chlorpyrifos. 3.3. Lipid content There was no clear change in the lipid content of the animals in any of the sediments during the 240 h study. The averages of the determined lipid percentages in each sediment through the experiments were used in the calculation of BSAFs. The lipid percentages (of fresh weight) were 0.9970.10% in S1, 0.9870.11% in S2, 1.2070.19% in S3 and 1.0270.17% in S4, as opposed to 1.1270.19% in the originating culture. 3.4. Egestion rate Sediment S3 was consumed in the greatest quantities by the animals (25–28 kg (dw) kg1 (fw)), followed by the

Table 4 Bioaccumulation factors (BAFs) and biota-sediment accumulation factors (BSAFs) for L. variegatus exposed to atrazine and chlorpyrifos in sediments S1–S4

Atrazine, S1 Atrazine, S2 Atrazine, S3 Atrazine, S4 Chlorpyrifos, Chlorpyrifos, Chlorpyrifos, Chlorpyrifos,

S1 S2 S3 S4

BAF (ks/ke)

BAF (Ca/Cs)

BSAF

0.186 0.548 3.856 19.562 3.071 28.077 25.337 48.860

0.128 0.276 1.752 14.504 1.849 13.925 17.160 57.304

4.582 1.937 5.320 2.490 67.109 99.181 34.954 6.219

other sediments in variable order (4–11 kg (dw) kg1 (fw)). The pesticide concentrations used in this study were not expected to affect the feeding rates of the animals, and the total dry weights of fecal pellets per the weight of the animals did not differ between the atrazine and chlorpyrifos experiments. The quantity of the compound ingested by the animals on basis of the weight of their fecal pellets generally seems to easily account for the quantity of the compound accumulated in the animals by the end of the experiment. However, the amount of sediment S4 apparently consumed only accounted for 45% of the accumulated atrazine, and with chlorpyrifos, the consumption of sediments S2 and S4 only accounted for 29% and 12% of the accumulated compound, respectively. 3.5. Desorption experiments The desorption data are presented in Fig. 2. Desorption from sediments S1–S3 was moderately quick and extensive, about 50% of the sorbate shifting into water within 24 h and 80–90% within 100 h. Desorption was the most rapid from S2, followed by S1 and S3. At the end of the study, 88–94% of the sorbate had been desorbed from sediments S1–S3. Desorption from S4 was, however, much slower and less complete. The recoveries of chlorpyrifos were 96.5% for S1, 97.7% for S2, 92.2% for S3, and 78.2% for S4.

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The three-compartment first-order desorption model was a better fit than the two-compartment model to the desorption data in all sediments, although both models were good fits (r240.99 for every sediment). However, according to statistical comparisons (extra sum-of-squares F-test) of the fit Eq. (3), the more complicated threecompartment model was statistically superior only for S1 and S3. The desorption parameters of both models are presented in Table 5. The modelled desorption curves (according to the model that is in each case the statistically superior one) are presented in Fig. 2. 4. Discussion 4.1. Bioaccumulation Both atrazine and chlorpyrifos are bioaccumulated in L. variegatus tissues. The bioaccumulation of chlorpyrifos was clearly stronger, as expected on basis of its greater lipophilicity (Table 4, Fig. 1). The uptake clearance coefficients (ks) increased in the same order for both chemicals (S1oS2oS3oS4), those of chlorpyrifos being distinctly larger than those of atrazine. This order

Chlorpyrifos left in sediment (%)

100

S1

S2

S3

S4

80

60

40

20

0 100

0

200

300 Time (h)

400

500

600

Fig. 2. The observed and modelled desorption of chlorpyrifos from sediments S1–S4.

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corresponds with the decreasing organic content among the sediments. The elimination rate coefficients (ke) are of the same order of magnitude for both chemicals and largest in sediment S4, where the animals bioaccumulated the greatest amount of both compounds. Since the steady state was not reached in the course of these experiments (240 h), the BAF values calculated as ¼ 240h) ¼ 0h) C(t /C(t could be expected to be lower than the a s ones predicted by the first-order kinetic model (ks/ke), as shown in Table 4. However, for chlorpyrifos bioaccumulation in the most inorganic sediment S4, the steady state seems to have been almost reached at the end of the experiment (Fig. 1). The BAF values seem to drop and BSAFs to grow with the organic content of the sediment. In theory, the BSAF of a compound should be constant if the OC content of the sediment is the only factor affecting the sorption and bioavailability of the compound. In practice, the bioavailability of both atrazine and chlorpyrifos is probably affected by other sediment qualities, such as the maturity of the organic material and the particle size distribution, as well the water solubility of the compounds and differences in their sorption behavior. The variability of the obtained BSAF values likely reflects differences in the quality of organic material and possibly that of other chemicalsorbing materials as well. Similar BSAF variability has been found in the bioaccumulation to L. variegatus of three pesticides (Ma¨enpa¨a¨ et al., 2003) and two surfactants (Ma¨enpa¨a¨ and Kukkonen, 2006). Since the egestion rates of the animals indicate feeding rates which do not entirely explain the bioaccumulation of either compound in S4, bioaccumulation may partly have occurred through some other route. As relatively watersoluble compounds, both atrazine and chlorpyrifos were probably bioaccumulated across the body wall from interstitial water. The animals may also have fed selectively on sediment particles containing the highest concentrations of the compounds; since S4 consists overwhelmingly of sand, the animals probably did feed selectively on its small but more nutritious organic fraction. You et al. (2006) determined the bioaccumulation of chlorpyrifos to L. variegatus from two artificially

Table 5 Desorption parameters for the desorption of chlorpyrifos from sediments S1–S4 kr (h1)

Fs

Two-compartment model S1 0.77 (0.02) S2 0.89 (0.01) S3 0.79 (0.01) S4 0.54 (0.03)

0.040 0.049 0.052 0.021

(0.002) (0.001) (0.002) (0.003)

0.20 0.11 0.19 0.43

Three-compartment model S1 0.13 (0.02) S2 0.24 (0.41) S3 0.19 (0.03) S4 0.062 (0.039)

0.27 (0.05) 0.086 (0.059) 0.20 (0.03) 0.46 (0.67)

0.71 0.67 0.65 0.53

Fr

ks (102 h1)

Fvs

kvs (103 h1)

r2

(0.02) (0.01) (0.01) (0.03)

2.3 (0.5) 1.1 (0.2) 1.1 (0.2) 0.73 (0.22)

N/A N/A N/A N/A

N/A N/A N/A N/A

0.9989 0.9996 0.9989 0.9982

(0.01) (0.40) (0.03) (0.04)

3.2 4.0 3.8 1.8

0.16 0.10 0.16 0.41

1.5 (0.2) 0.85 (0.27) 0.65 (0.09) 0.62 (0.25)

0.9999 0.9997 0.9999 0.9985

(0.1) (1.0) (0.2) (0.3)

(0.01) (0.01) (0.00) (0.04)

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contaminated sediments with OC contents comparable to those used in this experiment (1.31% and 7.85%) and comparable chlorpyrifos concentrations (18.6 and 3.5 mg kg1 OC, respectively). BSAF values at 14 d were 6.54 and 4.74, respectively. The BSAF value for the less organic sediment was therefore on the same level as that modelled for our most inorganic sediment (S4), while that for the sediment with the higher organic content was much lower than expected on basis of the high BSAF values we modelled for sediments S1 and S2 (Table 4). Differences in the quality of organic material may explain this. Since both pesticides were used in a radiolabelled form and analyzed with a LSC, the determined concentrations may consist partly of degradation products of the chemical after either metabolism in the animal or degradation in the sediment or the water column, although a cold fluorescent light source was used in order to prevent photodegradation. The results of our extractions of sediment samples taken at the beginning and the end of the study suggest that the extractability of chlorpyrifos in the sediments and therefore its concentrations as the parent compound did not diminish in the course of the 10-d study. You et al. (2006) also found 20% of chlorpyrifos in sediment to degrade over 28 d in a comparable study, which suggests that degradation in sediment over 10 d is not extensive. The extractability of atrazine, however, was notably lower at the end of our study in most sediments, which may have been due to either degradation or tight sorption to sediment particles in the course of the study. The concentrations of both pesticides determined in the animals in the course of our study are likely to represent the parent compound and their unexcreted and therefore relatively close metabolites. There is little literature available on the subject of metabolism of xenobiotics by L. variegatus, although Verrengia Guerrero et al. (2002) found the animal to fail to metabolize a selection of chlorophenols, PAHs, and pesticides in the course of an acute 48 h exposure period. 4.2. Desorption The desorption of chlorpyrifos was relatively slow as compared to that of the herbicide pendimethalin (log Kow 5.18) from the same sediments (Jantunen and Kukkonen, 2003), as well as the chlorpyrifos desorption measured with the same Tenax method from two sediments by You et al. (2006). Around 50–80% of the pendimethalin content of the sediments shifted into the water phase within the first 2–4 h, while 50% of the chlorpyrifos content in sediments S1–S3 desorbed in about 24 h. The shorter contact time (2 d) used in the pendimethalin experiment as opposed to 5 d in this experiment may explain some of this difference. However, You et al. observed quicker and more extensive chlorpyrifos desorption from their two sediments even after a 7 d contact time. The desorption behavior of chlorpyrifos from S4 (Fig. 2) was strange: desorption was notably slower and resulted in a higher desorption-resistant fraction than from the other,

more organic sediments. In an earlier experiment, the desorption of pendimethalin from this highly inorganic sediment was rapid and extensive (Jantunen and Kukkonen, 2003). The recovery of chlorpyrifos from S4 was also relatively low (78.2%) as compared to the other sediments. This, however, cannot account for the apparent slowness of the sorption, which indicates tight binding to some fraction of this sediment. There is little literature available on the sorption behavior of chlorpyrifos, although strong and irreversible sorption to humic acids, as well as strong sorption with a large adsorption–desorption hysteresis to some smectites have been reported (Wu and Laird, 2004; Van Emmerik et al., 2007). Since S3 and S4, the most inorganic sediments, showed the most incomplete desorption, chlorpyrifos would seem to desorb relatively well from organic material and more slowly from at least some type of inorganic material (sand, clay); however, in the study by You et al., the desorption of chlorpyrifos during the first 6 h was clearly more extensive from the sediment with the lower organic content. In our experiment, sediments S3 and S4 also displayed the highest BC/TOC contents (11% and 20%, respectively); however, even though the pore-water concentrations of chlorpyrifos were fairly low (0.18 mg L1 in S3, 3.0 mg L1 in S4), which would favor adsorption to BC, the high BSAF values of chlorpyrifos (24 in S3, 7.3 in S4) do not indicate that the organic content had a particularly strong chemical capacity for chlorpyrifos in any of the sediments. The clay fraction (o2 mm) of S4 has not been quantified but is likely to be low on basis of the coarse consistency of the sediment; however, it is possible that some of the sorbed chlorpyrifos was tightly bound to the clay fraction. The sediment with a low organic content used by You et al., originally a soil sample turned into sediment for experimental purposes by adding water, contained 14% sand, 70% silt and 16% clay, and therefore differed in grain-size composition from our sediment S4, which contained 63% sand and 47% smaller particles (Ma¨enpa¨a¨ et al., 2003). It is likely that the mineral composition of the inorganic materials of these sediments was also different. Such factors may explain the clearly slower desorption of chlorpyrifos from the more inorganic sediment S4. Since both pesticides were used in a radiolabelled form and analyzed with a LSC, which accounts for 14C rather than the compounds themselves, the determined concentrations may consist partly of degradation products of the chemical. However, the experiment was conducted under cold fluorescent light in order to prevent photodegradation, and a biocide was used in the desorption experiment in order to prevent biological activity, including biotransformation. Under the circumstances, neither pesticide is likely to have undergone extensive degradation. 4.3. The relationship of sediment qualities, sorption parameters, and bioaccumulation parameters No strong and significant correlations were found between any sediment qualities or sorption parameters

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and the bioaccumulation parameters, probably due to the small size of the dataset and the uneven distribution of most sediment qualities of interest among the sediments. As expected, the BAFs of both pesticides were negatively correlated with the KOC (the stronger correlation being with BAF ¼ ks/ke, which was 0.52 for atrazine and 0.88 for chlorpyrifos), but this correlation was not significant. In earlier studies with larger datasets, the rapid desorption rate of the nonplanar compound 2,4,5,20 ,40 ,50 hexachlorobiphenyl (HCBP) correlated negatively with the organic content of the sediment (Kukkonen et al., 2003) and the rapidly desorbed fraction of several chemicals (HCBP, benzo[a]pyrene (BaP), pyrene (PY) and 3,4,30 ,40 tetrachlorobiphenyl (TCBP)) was a good predictor of bioavailability to L. variegatus (Kukkonen et al., 2004). In this study, the rapid desorption rate of chlorpyrifos according to either the two- or three-compartment model did not correlate with the organic carbon content of the sediment, and while the rapidly desorbed fraction according to either model correlated positively with BSAF, the correlations (r2 ¼ 0.90 and 0.76 for Fr according to the two- and three-compartment desorption models, respectively) were not significant. While atrazine is nearly planar (Cornelissen et al., 2005b), chlorpyrifos cannot adopt a planar shape due to the phosphorothionate structure it contains. Kukkonen et al. (2005) found the BSAF values of a nonplanar compound (HCBP) to be consistent with EqP theory and to average 2.87 for L. variegatus in different sediments, while the BSAF values of three planar compounds (BaP, PY and TCBP) averaged below 1. The BSAF values acquired in this study for the nonplanar compound chlorpyrifos, however, are hardly consistent (range 6.2–99.2). The greater variability in the organic content of the sediments used in this experiment, as compared to those used by Kukkonen et al. (OC range 0.4–4%), may account for this difference; on the other hand, the BSAF values of atrazine in the same sediments were much more consistent (range 1.9–5.3). In the study by Kukkonen et al., correcting the o1 BSAF values for the additional sorption capacity of black carbon (BC) in the sediments seemed to create BSAF estimates more consistent with EQP (41). In this experiment, however, no BSAF values remained below unity and seem to need such a correction, which may be explained by the somewhat lower BC fractions of the sediments used in this study (average 0.09%) than those of the sediments used by Kukkonen et al. (average 0.2%). Furthermore, although atrazine is nearly planar, its environmental adsorption to BC has been found to be barely stronger than sorption to OC at low aqueous concentrations (1 ng L1) (Cornelissen et al., 2005b). In this study, atrazine concentrations in interstitial water were on mg L1 level. Aqueous concentrations above the ng L1 level together with BC forming only a small fraction of the total organic carbon are expected to greatly diminish the significance of BC as a sorbent (Cornelissen et al., 2005a). It is likely that the relatively

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low Kow of atrazine limits both its sorption to organic matter and its bioaccumulation. 5. Conclusions Both chlorpyrifos and atrazine bioaccumulated from several different sediments to L. variegatus. There was an expected clear difference between the relatively watersoluble atrazine and the more lipophilic chlorpyrifos: the latter bioaccumulated much more strongly, indicating that its lipophilicity strongly affects its bioaccumulation. The wide range and high level of BSAF values obtained particularly for chlorpyrifos, however, seem to challenge the mere size of the organic fraction as the major factor determining the bioaccumulation potential of a chemical from a particular sediment. Other influential factors likely include the quality (age, source material) of the organic matter and, at least in sediments with a low organic content, possibly also some properties of the inorganic materials. The desorption behavior of chlorpyrifos also differed markedly from that of pendimethalin, which has a similar lipid solubility (Kow) and has previously been studied comparably (Ma¨enpa¨a¨ et al., 2003; Jantunen and Kukkonen, 2003); the desorption of chlorpyrifos was slow and incomplete from a highly inorganic sediment (S4, OC 0.13%) which consisted largely of sand. This seems to indicate that some inorganic fraction of the sediment can reduce the bioavailability of chlorpyrifos. It is unlikely that adsorption to BC significantly affected the bioavailability of either chemical, since the BC contents of the sediments used in this experiment are low and the obtained BSAF values were fairly high. Our dataset cannot point out any specific sediment or chemical characteristics other than Kow and organic content which particularly affects the availability of these chemicals. Acknowledgments The work was funded by the project ‘‘Evaluation of availability to biota for organic compounds ubiquitous in soils and sediments’’ (ABACUS, EU contract EVK1-200100094) and the Academy of Finland (projects 206071 and 214545). We thank two anonymous reviewers for their comments and suggestions which helped to improve the manuscript. The work was funded by the project ‘‘Evaluation of availability to biota for organic compounds ubiquitous in soils and sediments’’ (ABACUS, EU contract EVK1-200100094) and by the Academy of Finland (projects 206071 and 214545). References Bondarenko, S., Gan, J., 2004. Degradation and sorption of selected organophosphate and carbamate insecticides in urban stream sediments. Environ. Toxicol. Chem. 23, 1809–1814.

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