Comparison of Earthworm Bioaccumulation between Readily

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Environ. Sci. Technol. 2010, 44, 323–328

Comparison of Earthworm Bioaccumulation between Readily Desorbable and Desorption-Resistant Naphthalene: Implications for Biouptake Routes YACHAO QI AND WEI CHEN* College of Environmental Science and Engineering/Tianjin Key Laboratory of Environmental Remediation and Pollution Control/Ministry of Education Key Laboratory of Pollution Processes and Environmental Criteria, Nankai University, Wei Jin Road 94, Tianjin 300071, China

Received September 23, 2009. Revised manuscript received November 18, 2009. Accepted November 18, 2009.

The bioaccumulation of readily desorbable naphthalene and that of desorption-resistant naphthalene in earthworms were compared to examine the effect of desorption resistance on bioavailability of hydrophobic organic contaminants in soil. A series of naphthalene-contaminated soil samples representing different degrees of desorption resistance were prepared using a batch sorption-repetitive desorption approach, and bioaccumulation of these samples was evaluated using earthworms (Eisenia fetida). Soil samples representing high-degree of desorption resistance exhibited considerably lower bioavailability, as indicated by the lower body burden (naphthalene concentration in worm tissue) at a given sorbed-phase naphthalene concentration. Moreover, the body burden of the highly desorption-resistant samples exhibited a significantly different pore-water dependency than that of the readily desorbable samples, indicating that different biouptake mechanisms are likely controlling readily desorbable contaminants and desorption-resistant contaminants. We propose that for readily desorbable contaminants, the primary biouptake route is the pore-water uptake, but for desorption-resistant contaminants enhanced uptake from ingested soil particles can also be important. The surfactant-like substances in worm gut fluids and physical stress such as abrasion are the likely causes for the enhanced release of desorption-resistant contaminants in worm guts. The difference in bioavailability betweenreadilydesorbableanddesorption-resistantcontaminants needs to be taken into account in risk assessment practices.

Introduction Sorption and desorption are very important processes controlling bioavailability of organic contaminants in soil environments (1). While bioavailability of hydrophobic organic contaminants such as polycyclic aromatic hydrocarbons (PAHs) is affected by many factors (e.g., physicochemical properties of the contaminants, types of organisms, environmental conditions, etc.), it is often assumed that only contaminants that can be released from soil/sediment are available to biological receptors (2-4). It is argued that even * Corresponding author phone/fax: 86-22-66229516; e-mail: [email protected]. 10.1021/es902899n

 2010 American Chemical Society

Published on Web 12/10/2009

contaminants that are ingested do not result in uptake or risk unless they can partition out of the sorbed-phase (5). Numerous laboratory and field studies have shown that only a fraction of contaminants sorbed in soil or sediment is readily desorbable, whereas the remaining fraction can be highly desorption-resistant. Desorption of the readily desorbable contaminants can be quantified with the general sorption/desorption models, but release of the desorptionresistant contaminants is hysteretic and does not follow the conventional sorption/desorption models; this phenomenon is typically referred to as desorption resistance, desorption hysteresis, slow desorption, or sequestration (6-8). Typical observations include lower equilibrium pore-water concentrations compared with the predicted values based on conventional sorption models, reduced extractability of aged soil samples, etc. (9, 10). While the mechanisms controlling desorption resistance have been a controversial topic for many years (and terms such as “readily desorbable” and “desorption-resistant” are largely operationally defined), it can be anticipated that readily desorbable contaminants and desorption-resistant contaminants exhibit considerably different bioavailability. Kraaij et al. (11) compared the accumulation of freshly spiked PAHs and PAHs in field-contaminated sediment to marine amphipods (Corophium volutator), and observed a good correlation between PAH accumulation and the rapidly desorbing fraction of soil-bound PAHs. They proposed that the difference in desorption behavior can largely explain the difference in bioaccumulation. Similarly, Lamoureux and Brownawell (12) proposed that in some sedimentary regimes, the bioavailability of lower molecular-weight compounds may be limited by slow desorption from a resistant phase. Reduced bioavailability to microorganisms has also been linked to aging or sequestration of sorbed contaminants (13, 14). More recently, Lu et al. (15) reported that compared with the reversibly sorbed phenanthrene, the desorptionresistant phenanthrene showed significantly reduced bioavailability to deposit-feeding tubificid oligochaete (Ilyodrilus templetoni). They proposed that pore-water concentration is a better indicator for the benthic worm bioaccumulation potential of sediment-associated organic contaminants (15, 16). In this study, we attempted to systematically compare the bioavailability of readily desorbable and desorptionresistant organic contaminants. A series of naphthalenecontaminated soil samples were preparedsby conducting a one-to-two-step sorption followed with one or multiple steps of repetitive desorptionsto represent different degrees of desorption resistance, and the bioavailability of sorbed naphthalene was evaluated by characterizing earthworm bioaccumulation using Eisenia fetida. (The rationale for selecting naphthalene as the model contaminant is discussed in the Supporting Information (SI).) Bioaccumulation of different samples was compared to establish the correlation with the degree of desorption resistance. The likely predominant biouptake mechanisms for the readily desorbable and desorption-resistant contaminants are proposed. Implications for risk assessment practices are discussed.

Materials and Methods Materials and Chemicals. A typical Chinese soils“black soil”swas used in this study. Soil collected from the surface layer (20 cm) was air-dried at room temperature, ground, and passed through a 2 mm sieve. The soil was slightly basic (pH 7.8) and contained 35% clay, 44% silt, and 21% sand. VOL. 44, NO. 1, 2010 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 1. Properties of Contaminated Soil Samples and Measured Body Burden (Cworm) and Biota-Soil Accumulation Factor (BSAF) Valuesa sample no.a S1 S2 S3 S4 S5 S6 S7 D1 D2 D3 D4 D5 D6 D7 D8

preparation procedures 2 2 1 1 1 1 1 2 2 1 1 1 1 1 1

sorption sorption sorption sorption sorption sorption sorption sorption sorption sorption sorption sorption sorption sorption sorption

f f f f f f f f

1 2 3 4 5 6 7 8

desorption desorption desorption desorption desorption desorption desorption desorption

q (mg/kg)

Cw (mg/L)

log KOC

78.1 59.1 39.1 11.1 7.90 4.30 0.780 112 87.5 44.3 16.3 13.1 12.0 5.60 5.50

5.99 3.78 2.68 0.645 0.425 0.278 0.0390 5.54 3.47 2.02 0.511 0.327 0.0568 0.0148 0.00490

2.79 2.87 2.84 2.91 2.94 2.87 2.98 2.98 3.08 3.02 3.18 3.28 4.00 4.26 4.73

HI

b

0.0090 0.26 0.094 0.60 1.0 9.5 17 48

Cworm

c

(mg/kg)

53.7 ( 5.6 36.4 ( 0.5 22.0 ( 0.5 4.24 ( 0.07 3.91 ( 0.34 1.99 ( 0.02 0.421 ( 0.041 66.2 ( 5.7 44.3 ( 2.1 21.4 ( 0.4 3.71 ( 0.07 2.99 ( 0.15 2.67 ( 0.05 1.10 ( 0.08 0.883( 0.091

BSAF

c

0.11( 0.03 0.095 ( 0.012 0.088 ( 0.017 0.061 ( 0.018 0.078 ( 0.023 0.076 ( 0.028 0.082 ( 0.012 0.092 ( 0.020 0.078 ( 0.010 0.076 ( 0.015 0.037 ( 0.011 0.036 ( 0.011 0.037 ( 0.006 0.031 ( 0.009 0.026 ( 0.009

a q, Cw, and KOC are sorbed-phase concentration, aqueous-phase concentration, and organic carbon-normalized distribution coefficient. a Letter S denotes “sorption”; letter D denotes “desorption”. b Hysteresis index. c Average of triplicate samples.

The fractional organic carbon (fOC) of the soil was 0.021. The soil did not contain detectable quantity of PAHs. Eisenia fetida was used as the test species. Eisenia fetida is a typical species of soil-dwelling earthworm, and is widely used in eco-toxicity testing (8, 17). Mature worms were obtained from Tianjin Biological Supply Co. (Tianjin, China) and were cultured in the laboratory. The average wet weight of the worms used in the bioaccumulation experiments was 0.29 ( 0.02 g each. Naphthalene was purchased from Sigma Chemical (St. Louis, MO). Stock solutions of naphthalene were prepared in methanol. All organic solvents were analytical grade or higher. Preparation of Contaminated Soil Samples. A total of seven naphthalene-contaminated soil samples were prepared via one-step or two-step sorption (these samples are named as S1 to S7, and are referred to as the “sorption group”), and a total of eight samples were prepared by one-or-two-step sorption followed by one or multiple steps of desorption (these samples are named as D1 to D8, and are referred to as the “desorption group”). The experimental protocols and the properties of these soil samples are summarized in Table 1. These samples represented different degrees of desorption resistance, and were used in bioaccumulation experiments immediately after preparation. To minimize the potential loss of naphthalene due to biodegradation over extended experimental duration, an electrolyte solution containing 0.01 M NaN3 was used in the sorption and repetitive desorption (until the last two desorption steps) for samples D3-D8. Sorption (i.e., soil contamination) was carried out using a batch sorption approach. First, a naphthalene aqueousphase solution was transferred to a 2.5-L volumetric flask containing 550 g of dry soil to initiate sorption. The container was sealed, and the soil suspension was continuously mixed with a magnetic stirrer until sorption equilibrium was reached. The potential loss of naphthalene to headspace was negligible (see SI). Afterward, the suspension was allowed to settle until the solution was clear by sight, and aliquots of the supernatant were withdrawn, centrifuged at 3000g for 30 min, and the aqueous-phase concentrations were measured. If another sorption was needed, a portion of the supernatant was removed and fresh naphthalene solution was added. Then, the procedures mentioned above were repeated. Finally, the soil slurry was centrifuged at 3000g to remove excess water before being used in bioaccumulation experiments. Repetitive desorption was conducted by successively replacing a portion of the supernatant with contaminant324

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free electrolyte solution. The time allowed for each repetitive desorption step was 5 days. (Apparent desorption equilibrium was reached in 3 days; refer to SI Figure S1). At the end of each desorption step, the soil suspension was allowed to settle until the solution was clear by sight. Aliquots of the supernatant were withdrawn, centrifuged at 3000g for 30 min, and the aqueous-phase concentrations were analyzed. Soil fOC value changed little with repetitive desorption (SI Table S1). Finally, the soil slurry was centrifuged at 3000g to remove excess water before being used in bioaccumulation experiments. Sorbed-phase concentrations were determined using a previously developed method (18). About 1 g wet soil was transferred to a 20 mL glass vial, and methanol-water solution (4:1 by volume) was added to leave approximately 1 mL of headspace. The vial was crimp-sealed and horizontally shaken in a water bath at 85 °C for 8 h. Afterward, the vial was centrifuged at 3000g for 30 min, and the supernatant was withdrawn for analysis. Extraction efficiencies determined in quality control experiments using spiked soil samples (with sorption time from 2-20 days) were 95-102%. Bioaccumulation Experiments. Bioaccumulation experiments were conducted using a method similar to that reported in the literature (14, 17). In a typical experiment, 10 worms were placed in a 500 mL glass beaker containing 150 g (dry weight) of naphthalene-contaminated soil. (One day before an experiment, worms of similar size were gently removed from the culture and held in Petri dishes until the initiation of the experiment.) The soil was adjusted to 45-55% moisture content with deionized water before the worms were added, and the beakers were covered with Saran wrap bearing holes for air entry. The beakers were kept in an incubator with a constant temperature of 25 °C for 14 days to ensure that steady state was reached (preliminary tests showed that accumulation of naphthalene by Eisenia fetida reached an apparent steady state within 7 days; see SI Figure S2). Triplicate samples were conducted for each soil sample. At the end of a bioaccumulation experiment, the worms were carefully removed, rinsed, and allowed to purge their gut contents for 24 h on moistened filter paper. All worms were active after the 14 day period. Then, the worms from each individual sample were weighed. (The change of worm mass before and after the bioaccumulation experiments was negligible.) Afterward, two worms from each beaker were taken for lipid analysis. The remaining worms were used to determine naphthalene concentration in worm tissue.

FIGURE 1. Comparison of the q-Cw correlation between soil samples prepared via sorption only (O) and samples prepared via sorption and repetitive desorption (9). The dashed line is plotted with the linear sorption model and solid line is plotted with a dual-equilibrium desorption (DED) model by Chen et al. (35). A method modified from the U.S. Environmental Protection Agency method 3550b was used to extract naphthalene from worm tissue. First, the worms were ground with approximately 10 g of anhydrous Na2SO4 with a mortar and pestle. Then, the tissue was transferred to a 40 mL crimpsealed vial and acetonitrile-water solution (4:1 by volume) was added to leave approximately 1 mL of headspace. The vial was sonicated for 30 min and was centrifuged at 3000g for 20 min. Finally, the extract was concentrated and cleaned with an OASIS HLB 3 cm3 solid-phase extraction column (Waters, Milford, MA) and analyzed for naphthalene concentration. Extraction efficiencies determined in quality control experiments using spiked tissue ranged from 95 to 101% (SI Table S2). Worm lipid content was determined using the method developed by Lu et al. (15). Two worms were transferred to a preweighed 25 mL centrifuge tube, and 10 mL of a methanol-chloroform cosolvent (1:1 by volume) was added. The mixture was sonicated for 1 min and allowed to equilibrate for 4 h. Then, the tube was centrifuged, and the supernatant was withdrawn and transferred to a new tube, and 10 mL of the methanol-chloroform cosolvent was added to do another extraction. The supernatant from the two extractions was combined and equilibrated with 5 mL of tap water to remove tissue protein. The extract was then dried at 50 °C and weighed to obtain the mass of lipid. The extraction efficiency of this procedure, determined by extracting a cholesterol standard (with 10 mg/mL lipid content), was 98%. The average lipid content of the worms used in this study was 12.9% (dry-weight based); the lipid content was similar among the worms and changed little before and after the bioaccumulation experiments. Analytical Methods. The concentrations of naphthalene in the aqueous-phase solution and in cosolvent were analyzed with a Waters 1100 high-performance liquid chromatography (HPLC) equipped with a binary HPLC pump 1525 (Waters, Milford, MA), a symmetry reversed phase C18 column (4.6 × 150 mm), and a Waters 2475 fluorescence detector. The mobile phase consisted of 85% acetonitrile and 15% water. The potential effect of colloids in the solution on the accuracy of dissolve-phase concentration measurement was evaluated and was found to be negligible (see SI).

Results and Discussion Characterization of Soil Samples. Figure 1 shows the q-Cw correlation (i.e., correlation between the sorbed-phase

FIGURE 2. Change of naphthalene accumulation in worm tissue (i.e., body burden) with naphthalene concentration in the soil. Symbol “O” represents soil samples S1-S7; Symbol “9” represents soil samples D1-D3; and Symbol “2” represents soil samples D4-D8. Error bars indicate standard deviations of triplicate samples. concentration and pore-water concentration) for the 15 contaminated soil samples. For the “sorption group” (samples S1-S7), the q-Cw correlation agrees reasonably well with the linear sorption model: q ) (KOC · fOC) · Cw, where q (mg/kg) and Cw (mg/L) are equilibrium concentrations of naphthalene in the soil and in the aqueous-phase solution, respectively, and KOC (L/kg-OC) is the organic carbon-normalized distribution coefficient. This indicates that sorption of naphthalene to the soil was controlled largely by hydrophobic partitioning to soil organic matter. For the “desorption group” (samples D1-D8), the q-Cw correlation deviates increasingly from the linear model with the number of repetitive desorption steps used in soil preparation (Table 1). The increasing desorption-resistance from sample D1 to sample D8 can also been seen from their apparent KOC values (Table 1): KOC increases gradually with the number of repetitive desorption, and for samples D6-D8 the apparent KOC values are over 1 order of magnitude higher than the KOC values of the “sorption group”. A hysteresis index value was calculated for each of the samples D1-D8. The hysteresis index is defined as (19): HI ) (qed - qes)/qes, where qed is a sorbed-phase concentration observed in the desorption experiment that is in equilibrium with an aqueous-phase concentration Ce; and qse is the sorbedphase concentration calculated from Ce assuming that desorption was reversible. If desorption is completely reversible, HI is equal to 0; the higher the HI value, the more desorption-resistant a sample is. The increase of the HI value with the number of desorption (Table 1) indicates that the mass fraction of the desorption-resistant naphthalene increased with repetitive desorption, and after multiple repetitive desorption (i.e., when the q value reached the plateau region of the desorption curve in Figure 1), a majority of the readily desorbable naphthalene was removed and the total sorbed mass was primarily associated with the mass of the desorption-resistant naphthalene (and the HI values became very high). Based on the HI values, samples D1-D8 should represent a wide range of degree of desorption resistance. Bioaccumulation of Readily Desorbable and DesorptionResistant Naphthalene. In Figure 2 the naphthalene concentration in worm tissue, Cworm (mg-naphthalene/kg-worm on a dry-weight base) (often referred to as body burden, i.e., mass of contaminant accumulated per unit mass of worm), is plotted against the sorbed-phase concentration (q) of naphthalene for different soil samples. For both soil samples prepared by one to two steps of sorption (samples S1-S7) and the samples prepared by sorption followed with repetitive desorption (samples D1-D8), the naphthalene accumulation VOL. 44, NO. 1, 2010 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 3. Correlation between biota-soil accumulation factor (BSAF) and the degree of desorption resistance indicated by the measured organic carbon-normalized distribution coefficient (KOC). Symbol “O” represents soil samples S1-S7; Symbol “9” represents soil samples D1-D3; and Symbol “2” represents soil samples D4-D8. Error bars indicate standard deviations of triplicate samples.

FIGURE 4. Correlation between naphthalene concentration in worm tissue and concentration in pore-water. Symbol “O” represents soil samples S1-S7; Symbol “9” represents soil samples D1-D3; and Symbol “2” represents soil samples D4-D8. Error bars indicate standard deviations of triplicate samples.

in worm tissue increased with the soil concentration (and good equilibrium partitioning correlations were observed; see equations in Figure 2). Nonetheless, the ratio of Cworm to q for the “desorption group” (samples D1-D8) is significantly lower than that of the “sorption group” (samples S1-S7) (P < 0.05, based on Kruskal-Wallis test of nonparametric analysis). This reduced bioaccumulation was particularly significant for the soils samples that exhibited higher degree of desorption resistance (samples D4-D8), the ratio of Cworm to q for this group is approximately 3 times lower than that for the “sorption group” (Figure 2). Bioaccumulation is the result of the influx into the organism on the one hand and the elimination by excretion of faeces and metabolism on the other hand. Thus, the lower bioaccumulation of desorption-resistant naphthalene was most likely due to a lower influx into the worm bodies. The difference between the bioavailability of readily desorbable naphthalene and that of desorption-resistant naphthalene can also be understood by comparing the biotasoil accumulation factor (BSAF) values of different soil samples (Table 1). BSAF is defined as (20)

fall into two distinct groups: the first group includes soil samples containing significant mass fraction of readily desorbable naphthalene (S1-S7, D1-D3), and the second group includes soil samples containing predominantly desorption-resistant naphthalene (D4-D8). The BSAF values of the second group are significantly greater than zero but over 50% lower than those of the first group. Interestingly, even though the BSAF values of the highly desorptionresistant naphthalene are significantly lower, among these soil samples (D4-D8) the BSAF value changed little. Similar observation has also been reported by Lu et al. (15), in that the BSAF values of reversibly sorbed phenanthrene were approximately 2 times higher than the values of the desorption-resistant phenanthrene, but the BSAF values among each group appeared to be similar. Thus, it might be reasonable to assume that up to a certain degree of desorption resistance, further increase of desorption resistance would not result in further reduction of bioavailability. Mechanisms for Biouptake. The bioaccumulation results mentioned above clearly indicate that readily desorbable organic contaminants and desorption-resistant organic contaminants exhibit significantly different bioavailability. A possible explanation is that only the desorbable contaminants are available for worm uptake and thus, bioaccumulation is better related to the pore-water concentration. (An analogy is that the worm body might be considered as a passive sampler-influx into the worm body depends on how much sorbed contaminants can be released into pore-water.) To test this hypothesis, the Cworm values of different soil samples were compared with the pore-water concentrations (Figure 4). Interestingly, Figure 4 clearly shows that even though for all the soil samples Cworm increases with porewater concentration, the three highly desorption-resistant samples (D6-D8) exhibit completely different pore-waterdependency than the other samples. For samples S1-S7 and D1-D5, Cworm shows near linear correlation with pore-water concentration (see equation in Figure 4). However, for samples D6-D8, the value of Cworm is markedly higher than what would be calculated based on the Cworm-Cw correlation derived from the other samples. This seems to indicate that for highly desorption-resistant samples, uptake from porewater alone cannot account for the total bioaccumulation, and other uptake route(s) could also make significant contributions to overall biouptake. In SI Table S3, the contribution of pore-water uptake to the overall bioaccumulation was estimated for samples D6-D8, based on the

BSAF )

Cworm /flipid q/fOC

(1)

where flipid is the lipid content of the worm. For hydrophobic and lipophilic compounds such as PAHs, BSAF is probably a better indicator for the bioaccumulation potential of a contaminant, because for such compounds the mass in the soil/sediment is mainly associated with the soil organic matter, and the mass in the worm tissue is mainly associated with the lipid. The BSAF values observed in this study are relatively low. Other researchers have also observed depressed bioaccumulation of more soluble hydrophobic organic compounds including naphthalene, and this has been attributed to the faster elimination rates for low-KOW compounds (21-23). Table 1 shows that the BSAF values of the more desorption-resistant samples (D4-D8) are substantially lower than the values of other samples (S1-S7 and D1-D3). The average BSAF value is 0.086 for samples S1-S7 and 0.083 for samples D1-D3, but is only 0.034 for samples D4-D8. To further examine the dependency of BSAF with the extent of desorption resistance, the BSAF values are plotted against the apparent KOC values of the soil samples (Table 1) in Figure 3. The figure shows that the BSAF values of the 15 soil samples 326

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linear Cworm-Cw correlation derived from samples S1-S7 and D1-D5, and the estimated values of percentage uptake from pore-water are only 22%, 14%, and 5.8% for samples D6, D7, and D8, respectively. Earthworms are able to take up organic chemicals in soil/ sediment both via dermal contact (uptake by the skin) and via direct ingestion of soil/sediment particles (24, 25). (For volatile compounds, uptake from soil gas is also possible.) The relative contributions of different uptake routes are controversial and unclear, but several studies have demonstrated the importance of the ingestion route (26-29). For example, Forbes et al. (27) postulated that the uptake of phenanthrene by marine infaunal polychaete (Capitella species I) from sediment was 20-30 times greater than that from porewater. Additionally, most earthworm species and benthic deposit-feeders are selective feeders. They can selectively ingest fine, nutritious particles (28, 29), and such selective feeding has been used to explain the elevated bioaccumulation levels compared with those estimated by the equilibrium partitioning models (20). It has been reported that the combined effects of physical, chemical, and biological processes inside the worm guts can markedly enhance the release of contaminants from ingested soil particles compared with the passive desorption of sorbed contaminants to pore-water (25, 30-33). First, the worm digestive fluids contain high concentrations of solubilizing agents, such as amino acids and surfactants, which can enhance desorption of both organics and metals (30-33). For example, Weston and Mayer (31) found that absorption of PAH solubilized by the gut fluids of the polychaete Arenicola brasiliensis was nearly 100%, showing the significance of desorption enhancement by gut fluids. Second, the gut wall consists of many small longitudinal folds and one larger fold, the typhlosole (32). Such unique physical structure not only results in a large inner surface suitable for the sorption of organics, but also facilitates the desorption of contaminants from the ingested soil/sediment due to the abrasive effect of the gut wall on soil particles. Some researchers even suggest that comminution of mineral particles occurred during the passage through the gut of an earthworm (32). Furthermore, Mayer et al. (33) also demonstrated enhanced diffusive mass transfer of PAHs through the digestive fluid of a benthic worm. Based on the discussions above, it might be reasonable to postulate that bioaccumulation of readily desorbable contaminants and bioaccumulation of desorption-resistant contaminants are controlled by different uptake routes. For readily desorbale contaminants, uptake from pore-water is likely the dominant route. For highly desorption-resistant contaminants, enhanced uptake from ingested soil particles might be a more dominant route because contaminant concentrations in pore-water are too low. Nonetheless, it is necessary to note that the biouptake processes involving soils/ sediments are very complicated. For example, natural soils likely contain a continuum of sorption sites each characterized by different sorption energies and sorption kinetics, and the degree of desorption resistance can be further complicated by slow diffusion (in soil organic matters or micropores of soil particles) as well as physical entrapment in soil matrix (6-8). Surely, either passive desorption into pore-water or enhanced release in gut fluids can be markedly different for different soil components. Moreover, bioaccumulation is a dynamic process affected by many factors such as desorption rates, various diffusion rates (e.g., diffusion out of soil, diffusion though worm skin, etc.), and potential metabolisms of contaminants in worm gut. Therefore, attention should also be paid to the kinetics aspects of bioaccumulation. For instance, it might be useful to compare the accumulation kinetics of the “live” worms with the “dead” passive samplers and to examine residual contaminants in faeces to better

understand the effects of digestive tract. Undoubtedly, many more studies that involve more chemicals of different physiochemical properties and different types of contaminated soils are needed to fully reveal the mechanisms controllingaccumulationofdesorption-resistantcontaminants. Implications for Risk Assessment. Findings of this study clearly demonstrate that readily desorbable and desorptionresistant contaminants exhibit different bioavailability. However, most of the current risk assessment practices are based on the assumption that the entire mass of a contaminant in soil is equally available to desorption or to biological receptors, and the effect of desorption resistance is often not considered. This approach could significantly overestimate the risk of soil/sediment-associated organic contaminants. Indeed, many laboratory and field tests have demonstrated that bulk soil and sediment concentrations yield a wide range of toxic response results (17, 34), and desorption resistance can be an important factor resulting in such discrepancies and needs to be taken into account in risk assessment practices. Moreover, findings of this study indicate that even though desorption-resistant contaminants exhibit considerably lower bioavailability to earthworm compared with readily desorbable contaminants (which was expected), simply applying the pore-water approach (i.e., assuming bioaccumulation is controlled by pore-water concentration) can underestimate the bioavailability of highly desorptionresistant contaminants. An important future research direction might be to derive general equilibrium partitioning correlations between biouptake and soil/pore-water concentrations for both the readily desorbable and desorptionresistant contaminants for typical organisms and environmental conditions. Such correlations would greatly facilitate risk assessment and ecological benchmark calculations.

Acknowledgments This project was supported by National Natural Science Foundation of China (Grants 20637030, 20407013), Ministry of Science and Technology (Grant 2009DFA91910), and China-U.S. Center for Environmental Remediation and Sustainable Development. We thank Professor Rong Ji, Nanjing University, for his input on earthworm digestion processes.

Supporting Information Available Table S1 shows the change of soil fOC value with repetitive desorption. Table S2 summarizes the extraction efficiency of naphthalene from spiked tissue. Table S3 shows the contribution of nonpore-water uptake to total uptake for samples D6-D8. Figure S1 shows the apparent desorption equilibrium times. Figure S2 shows the accumulation kinetics of naphthalene by Einsenia fetida. The rationale for selecting naphthalene as the model contaminant, an analysis of the potential loss to headspace during sample preparation, and an analysis of colloidal effect on aqueous-phase concentration measurement are included. This material is available free of charge via the Internet at http://pubs.acs.org.

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