1H NMR metabolomics of earthworm exposure to sub-lethal concentrations of phenanthrene in soil

Environmental Pollution 158 (2010) 2117e2123

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H NMR metabolomics of earthworm exposure to sub-lethal concentrations of phenanthrene in soil Sarah A.E. Brown, Jennifer R. McKelvie, Andre J. Simpson, Myrna J. Simpson* Department of Physical and Environmental Sciences, University of Toronto, 1265 Military Trail Toronto, Ontario, Canada M1C 1A4 1

H NMR metabolomics is used to directly monitor metabolic responses of Eisenia fetida after 48 h of exposure to sub-lethal concentrations of phenanthrene in soil.

a r t i c l e i n f o

a b s t r a c t

Article history: Received 1 September 2009 Received in revised form 14 February 2010 Accepted 28 February 2010

1 H NMR metabolomics was used to monitor earthworm responses to sub-lethal (50e1500 mg/kg) phenanthrene exposure in soil. Total phenanthrene was analyzed via soxhlet extraction, bioavailable phenanthrene was estimated by hydroxypropyl-b-cyclodextrin (HPCD) and 1-butanol extractions and sorption to soil was assessed by batch equilibration. Bioavailable phenanthrene (HPCD-extracted) comprised w65e97% of total phenanthrene added to the soil. Principal component analysis (PCA) showed differences in responses between exposed earthworms and controls after 48 h exposure. The metabolites that varied with exposure included amino acids (isoleucine, alanine and glutamine) and maltose. PLS models indicated that earthworm response is positively correlated to both total phenanthrene concentration and bioavailable (HPCD-extracted) phenanthrene in a freshly spiked, unaged soil. These results show that metabolomics is a powerful, direct technique that may be used to monitor contaminant bioavailability and toxicity of sub-lethal concentrations of contaminants in the environment. These initial findings warrant further metabolomic studies with aged contaminated soils. Ó 2010 Elsevier Ltd. All rights reserved.

Keywords: Eisenia fetida Metabonomics Metabolomics Bioavailability Earthworm responses

1. Introduction Soil contamination is typically assessed by exhaustive extraction methods (e.g. soxhlet extraction); however, the quantities measured do not necessarily correlate to the contaminant concentration available to soil biota (Reid et al., 2000; Doick et al., 2006; Swindell and Reid, 2006; Barthe and Pelletier, 2007; Rhodes et al., 2008). Contaminants in soils have varying degrees of extractability (Semple et al., 2003) and several soft chemical extractions have shown positive correlations with contaminant mineralization, suggesting that these extractions may be suitable indirect measures of chemical bioavailability (Semple et al., 2003; Dean and Scott, 2004; Doick et al., 2006; Barthe and Pelletier, 2007). Several lines of evidence exist which demonstrate that the levels of contamination (i.e.: concentration of the contaminants extracted from soil) are poor predictors of contaminant risk and toxicity (Kelsey and Alexander, 1997; Johnson et al., 2002; Fent, 2003; Dean and Scott, 2004; Bergknut et al., 2007). This suggests a need to develop direct tools for assessing contaminant

* Corresponding author. E-mail address: [email protected] (M.J. Simpson). 0269-7491/$ e see front matter Ó 2010 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2010.02.023

bioavailability and to further investigate the relationship between bioavailability and the exposure response of earthworms. Metabolomics has become a popular technique for identifying the metabolic responses of earthworms to sub-lethal contaminant exposure and 1H nuclear magnetic resonance (NMR) has proven to be a powerful approach in several pioneering metabolomic studies (Warne et al., 2000; Bundy et al., 2001, 2002, 2004, 2007; Jones et al., 2008a; Guo et al., 2009; McKelvie et al., 2009; Rochfort et al., 2009). Environmental metabolomic studies using earthworms predominantly employ 1H NMR (Warne et al., 2000; Bundy et al., 2001, 2002, 2004, 2007; Lin et al., 2006; Jones et al., 2008a; Guo et al., 2009; McKelvie et al., 2009; Rochfort et al., 2009), a fast, non-selective, and non-destructive technique that requires minimal sample preparation (Bundy et al., 2002; Dunn and Ellis, 2005; Lenz et al., 2005; Robertson, 2005; Lin et al., 2006; Lenz and Wilson, 2007; Jones et al., 2008a). Statistical methods such as principal component analysis (PCA) and/or partial least squaresdiscriminant analysis (PLS-DA) are often used to screen 1H NMR spectra to identify metabolite fluxes between exposed and unexposed earthworms (Warne et al., 2000; Bundy et al., 2001, 2002, 2004, 2007; Robertson, 2005; Jones et al., 2008a; Guo et al., 2009; McKelvie et al., 2009; Rochfort et al., 2009). 1H NMR spectra can be evaluated to identify novel metabolites formed due to contaminant exposure, or quantified to determine the extent to

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which endogenous metabolites increase or decrease (Lenz et al., 2005; Malz and Jancke, 2005; Lin et al., 2006). This approach has been used in contact tests (Warne et al., 2000; Bundy et al., 2002; McKelvie et al., 2009) and soil studies (Bundy et al., 2004, 2007; Jones et al., 2008a; Guo et al., 2009; Rochfort et al., 2009) with a range of environmentally problematic chemicals but to date, attempts to link earthworm responses to bioavailability have not been made. In this study we examine the response of Eisenia fetida, measured by 1H NMR metabolomics, after exposure to eight sublethal concentrations of phenanthrene in soil. E. fetida is the earthworm species recommended by the Organization for Economic Cooperation and Development (OECD) for use in toxicity tests (OECD, 1984) but has not been used in metabolomic studies to the same extent as other earthworm species (Warne et al., 2000; Bundy et al., 2001, 2002, 2004, 2007; Jones et al., 2008a; Guo et al., 2009; McKelvie et al., 2009). In addition, metabolomic studies have not yet investigated the relationship between phenanthrene bioavailability [estimated by hydroxypropyl-b-cyclodextrin (HPCD) extraction and 1-butanol extraction] and earthworm metabolomic responses in soil. We used two bioavailability proxies in this study (HPCD and 1-butanol) because while both these methods can provide an estimate of the bioavailable amount of PAH in soil (Reid et al., 2000; Liste and Alexander, 2002; Semple et al., 2003; Barthe and Pelletier, 2007), their extraction efficiency can vary with soil composition and aging (Reid et al., 2000; Swindell and Reid, 2006). However, it must be noted that we used an unaged soil in this study and thus the concentration of bioavailable phenanthrene in our freshly spiked soil may not be representative of bioavailable phenanthrene in an aged soil. Therefore, this study aims to test the feasibility of using earthworm metabolomics as an additional measure of bioavailability. The objective of this study is to characterize and quantify earthworm responses after exposure to a range of sub-lethal phenanthrene concentrations in soil using 1 H NMR metabolomics and evaluate the relationship between earthworm metabolomic responses, phenanthrene bioavailability (as measured by proxy) and total phenanthrene concentration in a freshly spiked soil. The overall goal is to investigate the potential of 1H NMR metabolomics to directly monitor contaminant bioavailability in soil and evaluate the associated risks of contaminant exposure. 2. Materials and methods 2.1. Soil spiking A sphagnum peat soil (Magic Products Inc., WI, USA) was used because it consistently maintains optimal earthworm health (Brown et al., 2008). Soil was sieved to remove particles >2 mm and analyzed for total organic carbon prior to use (Supplementary material, Section S1). The amount of phenanthrene sorbed to soil during the experiment relative to the amount in solution was measured [Supplementary material, Section S2 and Fig. S1; (Bonin and Simpson, 2007)]. The soil-water equilibrium distribution coefficient (Kd) was calculated from the slope of the sorption isotherm (Supplementary material, Fig. S1) and the organic carbon normalized sorption coefficient (Koc) was calculated by dividing the Kd value by the fraction of organic carbon (foc ¼ 0.266) [Supplementary material, Fig. S1; (Bonin and Simpson, 2007)]. Preliminary experiments revealed the range of sub-lethal phenanthrene concentrations (Supplementary material, Section S3 and Table S1). Therefore, eight spiked soils were prepared with phenanthrene concentrations 50e1500 mg/kg (dry wt) for use in the metabolomic study. Four controls included soil (Control 1), soil þ dichloromethane (DCM; Optima, Fisher Scientific, ON, Canada; Control 2), soil þ DCM þ earthworms (Control 3) and soil þ 250 mg/kg phenanthrene (Control 4). Control 1 and 2 were used to ensure that there was no phenanthrene contamination of the soil or solvent (DCM), Control 3 was used in the metabolomic study and Control 4 was used to identify whether earthworms enhance phenanthrene removal during the experiment. Twenty grams of soil in 1 L clear glass jars was spiked by adding 4 mL of 1e30 mg/mL phenanthrene (>96% purity, SigmaeAldrich, ON, Canada) in DCM (for phenanthrene-spiked soils and Control 4) or 4 mL DCM (Controls 2 and 3; Brinch et al., 2002). The soil was mixed thoroughly (using a stainless steel spatula) and after 5 min DCM was allowed to

evaporate for 16 h (Brinch et al., 2002). An additional 60 g of soil was added to each jar and thoroughly mixed. The soil was then wetted with distilled water (160 mL water/80 g soil) and the water was allowed to absorb for 24 h before earthworms were added. Soil water content was constant throughout the duration of the experiment (73.4%  0.5%, Supplementary material, Table S2). 2.2. Earthworm exposure and preparation for 1H NMR E. fetida earthworms (The Worm Factory, Kingston, ON, Canada) were maintained as described in Brown et al. (2008). Ten mature earthworms were added to each jar filled with phenanthrene-spiked soil and the control soil (Control 3), prepared as described in Section 2.1. Earthworms were kept in closed jars for 48 h at 21  C and in natural light to encourage burrowing in the soil (Eijsackers et al., 2001). This temperature is within the optimal range for E. fetida (20e29  C; Presley et al., 1996). The earthworms were then removed and depurated for 96 h on damp filter paper to purge their intestinal tract (Brown et al., 2008). Earthworms were weighed, flash frozen in liquid N2, lyophilized, homogenized using a 5 mm stainless steel spatula and finally extracted into 1 mL of a D2O-based (99.9% purity, Cambridge Isotope Laboratories, MA, USA) 0.2 M sodium phosphate buffer (NaH2PO4$2H2O; 99.3% purity, Fisher Chemical), adjusted to pD 7.4 with sodium deuteroxide [NaOD; 30% w/w in 99.5% D2O, Cambridge Isotope Laboratories, (Brown et al., 2008)]. This extraction method isolates the largest quantity of the widest variety of metabolites from E. fetida (Brown et al., 2008). The buffer contained 10 mg/L of 2, 2-dimethyl-2silapentane-5-sulfonate sodium salt (DSS, 97% purity, Sigma Aldrich) as an internal standard and 0.1% (w/v) sodium azide (NaN3, 99.5% purity, Sigma Aldrich) to minimize sample degradation (Bundy et al., 2001; Brown et al., 2008). DSS is used as an internal standard for aqueous samples and has been used in other studies where quantification from 1H NMR spectra was performed (Saude et al., 2006; Brown et al., 2008). The earthworm homogenate was mixed using a VX 100 vortexer (Labnet, NJ, USA), sonicated for 15 min using an FS60 sonicator (Fisher Scientific) and centrifuged at 12 000 rpm (9177 g) for 15 min on a Hanil Micro-12 centrifuge [Rose Scientific, AB, Canada (Brown et al., 2008)]. The pellet was discarded, the supernatant was recentrifuged for 5 min to remove residual particles and then transferred into 5 mm tubes (Norell Inc., NJ, USA) for 1H NMR (Brown et al., 2008). 2.3. Total phenanthrene concentration and bioavailable phenanthrene Total phenanthrene present in the soil after 48 h was extracted using a modified version of EPA Method 3540C (U.S., 1996). Soils were soxhlet extracted in triplicate. 5 g of soil, mixed with 5 g of sodium sulfate (Na2SO4; min 99.5% purity, Fisher Scientific) and placed in an extraction thimble, was soxhlet extracted for 18 h using 150 mL DCM. Extracts were concentrated by rotary evaporation, and then diluted with DCM to 10 mL. 2 mL aliquots were transferred to amber glass vials (Agilent Technologies Inc.) and the phenanthrene concentration was quantified by gas chromatographyemass spectrometry (Supplementary material, Section S4). Bioavailable phenanthrene was estimated by 1-butanol and HPCD extractions (Reid et al., 2000; Barthe and Pelletier, 2007). For 1-butanol extraction, 1.25 g of each soil was weighed into 50 mL Teflon centrifuge tubes and 10 mL of 1-butanol (>99.4% purity, Fisher Scientific) was added. The tubes were sealed, put on an orbital shaker for 20 h (Reid et al., 2000), then centrifuged at 2700 rpm (1150 g) for 20 min. The supernatant was decanted and filtered through Whatman GF/F glass microfibre filter paper (Fisher Scientific). 2 mL supernatant was transferred to amber glass vials and the phenanthrene concentrations were quantified by high performance liquid chromatography with UV detection (Supplementary material, Section S5). For the HPCD extraction, 1.25 g of each soil was weighed into 50 mL Teflon centrifuge tubes and 25 mL of 50 mM HPCD (97% purity, Acros Organics, ON, Canada) was added (Reid et al., 2000; Barthe and Pelletier, 2007). The tubes were sealed, put on an orbital shaker for 20 h (Reid et al., 2000), then centrifuged at 2700 rpm (1150 g) for 20 min. The supernatant was decanted and filtered through Whatman GF/F glass microfibre filter paper (Fisher Scientific). 10 mL supernatant was mixed with 10 mL methanol (Optima, Fisher Scientific) shaken and then sonicated for 1 h to breakdown the HPCD (Barthe and Pelletier, 2007). 2 mL of the methanol-supernatant mixture was transferred to amber glass vials and phenanthrene concentrations were quantified by high performance liquid chromatography with UV detection (Supplementary material, Section S5). 2.4.

1

H NMR metabolomics

1 H NMR spectra of earthworm extracts were acquired with a Bruker Avance 500 MHz spectrometer (Bruker BioSpin, Rheinstetten, Germany) with a 1H-BB-13C Triple Resonance Broadband Inverse probe fitted with an actively shielded Zgradient. 1H NMR experiments were performed using Presaturation Using Relaxation Gradients and Echoes (PURGE) water suppression and 128 scans, a recycle delay of 3 s, and 16 K time domain points (Simpson and Brown, 2005). Spectra were apodized through multiplication with an exponential decay corresponding to 0.3 Hz line broadening in the transformed spectrum and a zero filling factor of 2. All spectra were baseline corrected and manually phased, then calibrated to the methyl resonance signal of the internal standard (DSS, d ¼ 0.00 ppm; Saude et al., 2006). Using the AMIX Statistics package (Version 3.8.4, Bruker BioSpin) PCA was performed on

S.A.E. Brown et al. / Environmental Pollution 158 (2010) 2117e2123 the 1H NMR dataset from 0.50 to 10.00 ppm, excluding the methyl resonance signal of DSS at 0.00 ppm (all exposed earthworms and controls; PCA data shown in Supplementary material, Section S6). The 1H NMR spectra were divided into sections 0.02 ppm wide and the area under each segment was integrated, creating a bucket table (the X-matrix). The area from 4.70 to 4.90 ppm was excluded because this region includes small residual signals from HOD and H2O. The integration mode was set to sum of intensities and spectra were scaled to total intensity. PCA was performed at a confidence level of 95% and variances representing less than 1% were excluded. The mean PCA scores and associated standard error for all the exposed earthworm sets and controls were also calculated to determine trends with soil phenanthrene concentration (Fig. 1A). Loadings plots of the PCs were derived from PCA results to identify regions of the 1H NMR spectra that contribute to the observed variance (McKelvie et al., 2009). PLS was performed on the entire dataset using the AMIX statistics package (Version 3.8.4, Bruker BioSpin), to investigate the relationship between earthworm exposure response to both total and bioavailable phenanthrene. PLS analysis was conducted using the NIPALS PLS algorithm (Vendeginste et al., 1998). The R2X and R2Y values, which explained variation of X and Y, respectively (Wold et al., 2001; Eriksson et al., 2006), and the cross validated R2X and R2Y values [reported as Q2X and Q2Y, respectively (Baroni et al., 1993; Cramer, 1993; Wold and Eriksson, 1995; Lee et al., 2003; Mitra et al., 2010)] were calculated with AMIX (Version 3.8.4, Bruker BioSpin). The same bucket table was used as in PCA and Y-tables were created that identified the quantified total (soxhlet-extracted) and bioavailable (HPCD-extracted) phenanthrene concentration. The PLS models were cross validated using leave one out cross validation (See Supplementary material, Section S7 for details). The predictive power of the PLS models were also tested by creating models using only w80% of the dataset that was then used to predict the remaining data (See Supplementary material, Section S10 for details). Metabolites with signals corresponding to the regions of the 1H NMR spectra that contributed to the variance in the datasets were identified by comparison to 1H NMR spectra of metabolites previously identified in E. fetida (Gibb et al., 1997; Brown et al., 2008). The buckets corresponding to the metabolite signals that did not overlap with other identified metabolites (isoleucine, leucine, valine, alanine, glutamine, lysine and maltose) were quantified. The mean (n ¼ 10) bucket intensity and associated standard error were calculated for exposed and control earthworms

PC4 (1.7% variance)

A

2119

and the percent intensity change was calculated using the equation (IE  IC)/IC  100, where IC is the mean bucket intensity of the selected metabolite for control earthworms and IE is the corresponding mean bucket intensity for the exposed earthworm set. A two sample t-test (two tailed, assuming equal variance) was performed comparing each set of exposed earthworms to the control earthworm set, to identify if the change in bucket intensity was significant (p < 0.05 and/or p < 0.10; Ekman et al., 2008, 2009). Metabolite concentrations were also calculated from the 1H NMR spectra, see supplementary material for method (Section S8).

3. Results and discussion 3.1. Phenanthrene sorption and bioavailability Total soil phenanthrene concentration showed that degradation and/or volatilization of phenanthrene did not occur over 48 h (Supplementary material, Table S3). In addition, total phenanthrene in soxhlet extracts of control soil spiked with 250 mg/kg phenanthrene (Control 4) and soil spiked with 250 mg/kg phenanthrene that contained earthworms had comparable phenanthrene contents after 48 h. Sorption coefficients (distribution coefficient; Kd ¼ 1722  72 mL/g and organic carbon normalized distribution coefficient; Koc ¼ 6470  269 mL/g) were both within the range for soils studied previously (Bonin and Simpson, 2007). The concentration of phenanthrene sorbed to the spiked soil, estimated using Kd and the soil-water content, is high (w99.8%). 1-Butanol extracted the same concentration of phenanthrene as soxhlet extraction (data not shown). HPCD-extractable phenanthrene ranged from w65 to 97% of the total phenanthrene concentration determined from soxhlet extraction and increased with increasing total phenanthrene concentration (Supplementary material, Table S3 and

Control

0.08 0.04

High concentration 0.00 -0.04 Low concentration -0.08 -0.40

-0.20

0.00

0.20

0.40

0.60

PC1 (74.3% variance)

B

C

Leu/Lys Val Arg/Lys Ala Leu

Positive PC1 loadings

Glu/Gln

Overlapping sugars Leu and amino acids Leu/ Lys Val Lys Maltose Arg/ Lys

Positive PC4 loadings

Ile

LysGln

Negative PC1 loadings

Gln

Negative PC4 loadings Ile

Maltose Overlapping sugars and amino acids

10.00

7.50 5.00 2.50 Chemical shift (ppm)

Ala

0.00

10.00

7.50

5.00

2.50

0.00

Chemical shift (ppm)

Fig. 1. A) PCA scores plot of the mean (n ¼ 10) scores of the integrated 1H NMR spectra of E. fetida control earthworms (C) and earthworms exposed to soil spiked with 50 mg/kg (B), 75 mg/kg (6), 125 mg/kg (,), 250 mg/kg ( ), 500 mg/kg (>), 750 mg/kg (þ), 1000 mg/kg () and 1500 mg/kg (7) phenanthrene. The error bars represent standard error and the ovals indicate groups of earthworms exposed to phenanthrene at low (50e500 mg/kg) and high (750e1500 mg/kg) concentrations. B) The corresponding PC1 and C) PC4 loadings versus chemical shift. Spectral regions contributing to the observed variance are labeled with selected metabolites. Ile ¼ isoleucine, Leu ¼ leucine, Val ¼ valine, Ala ¼ alanine, Lys ¼ lysine, Arg ¼ arginine, Gln ¼ glutamine, Glu ¼ Glutamic acid. Two earthworms exposed to 1000 mg/kg phenanthrene did not survive exposure and their data was not used in calculations.

S.A.E. Brown et al. / Environmental Pollution 158 (2010) 2117e2123

H NMR metabolomics of earthworm responses

Aqueous extractions are frequently used in earthworm metabolomics (Bundy et al., 2001, 2004; McKelvie et al., 2009) and the major metabolites in a D2O-buffer extract of E. fetida have been previously identified (Brown et al., 2008). The mean 1H NMR spectra (Supplementary material, Fig. S3) of E. fetida exposed to soil spiked with phenanthrene contained the same major metabolites as unexposed (control) earthworms and neither novel metabolites nor phenanthrene were detected by visual inspection (Brown et al., 2008). PCA was used to identify whether phenanthrene-exposed earthworms differed from controls. This unsupervised multivariate statistics method is commonly used to discriminate between exposed and control earthworms (Warne et al., 2000; Bundy et al., 2001, 2002, 2004, 2007; Guo et al., 2009; McKelvie et al., 2009; Rochfort et al., 2009). The mean PCA scores for all the exposed and control earthworms were calculated to test concentrationdependence of responses (Warne et al., 2000). The PCA scores plots for the 1H NMR spectra of each exposed earthworm group versus controls and their corresponding loadings plots are shown and discussed in the Supplementary material (Section S6 and Figs. S4 and S5). The mean scores plot of PC1 versus PC4 for all the exposed earthworms and controls (Fig. 1A) shows that exposed earthworms clustered apart from controls and 76.0% of variance was explained by both PC1 (74.3%) and PC4 (1.7%). There is a clear response to phenanthrene exposure, earthworms exposed to 750e1500 mg/kg phenanthrene (high concentrations) separate from controls along PC1 and the average scores fall outside standard error. Earthworms exposed to 50e500 mg/kg phenanthrene (low concentrations) separate from controls along PC4 and the average scores fall outside standard error except for earthworms exposed to 75 mg/kg phenanthrene. Because separation from controls occurs along PC1 for earthworms exposed to high concentrations of phenanthrene, which explains a greater amount of variance than PC this indicates that earthworms exposed to high concentrations of phenanthrene have a greater response than those exposed to low concentrations of phenanthrene. These results suggest that earthworm responses are dependent on the exposure concentration. The PC1 loadings plot (Fig. 1B) showed that the spectral regions from w3.00e5.50 ppm and w0.50e2.50 ppm contribute to the variance in PC1, while in PC4 loadings (Fig. 1C) the loadings from w3.00e5.50 ppm are less intense than in the PC1 loadings. Therefore, with exposure to high phenanthrene concentrations (750e1500 mg/kg), variation in maltose and overlapping sugar and amino acid resonances as well as the amino acid resonances (Brown et al., 2008) were contributing to the separation from controls, but with exposure to low concentrations of phenanthrene, the variation in maltose and overlapping sugar and amino acid resonances are less important.

% intensity change

1

A

Isoleucine

p=0.02*

40.0 20.0 0.0 50

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125

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500

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Phenanthrene concentration (mg/kg soil) % intensity change

3.2.

The percent bucket intensity change was calculated for metabolites identified from the loadings plots that contributed to the variance in PC1 and PC4. The metabolites selected were those with signals that did not overlap with other identified metabolite signals (Gibb et al., 1997; Brown et al., 2008). There were no significant changes in leucine, valine or lysine levels for earthworms exposed to 50e1500 mg/kg phenanthrene. However, isoleucine increased in earthworms exposed to 50e1500 mg/kg phenanthrene-spiked soil (Fig. 2A) and the difference was significant (p < 0.05) for earthworms exposed to 1000 mg/kg phenanthrene compared to control earthworms. Alanine also increased in all exposed earthworm groups (Fig. 2B) and the difference was significant (p < 0.05 and/or p < 0.10) for earthworms exposed to 50, 75, 250, 750 and 1000 mg/ kg phenanthrene compared to controls. Glutamine increased

120.0 100.0

B

Alanine

p=0.03*

80.0 60.0 40.0 20.0 0.0

p=0.05+ p=0.06+

p=0.07+ p=0.08+

50

75

125

250

500

750

1000

1500

Phenanthrene concentration (mg/kg soil)

% intensity change

Fig. S2A). The varying degree of extractability between methods is likely because the soil was spiked over 48 h [i.e., the soil was not aged (Swindell and Reid, 2006)]. Therefore, HPCD-extractable phenanthrene was used here to estimate bioavailable phenanthrene concentrations. The bioavailable concentration (Supplementary material, Fig. S2B) was greater than the estimated soil pore water concentration, suggesting that both phenanthrene dissolved in soil pore water and sorbed phenanthrene were extracted by HPCD (Semple et al., 2004). The remaining nonextractable phenanthrene was deemed non-bioavailable (Supplementary material, Fig. S2B) and likely tightly sorbed to soil. The percentage of phenanthrene removed by HPCD extraction was comparable to that found in other studies using phenanthrenespiked soils with a similar percentage of organic carbon (Reid et al., 2000; Swindell and Reid, 2006; Rhodes et al., 2008).

30.0

C

Glutamine

p=0.004*

p=0.003* p=0.01*

20.0 10.0 0.0 -10.0 50

75

125

250

500

750

1000

1500

Phenanthrene concentration (mg/kg soil)

% intensity change

2120

20.0

D

Maltose

0.0 -20.0 -40.0 p=0.02*

-60.0 -80.0 50

75

125

250

500

750

p=0.01*

1000

p=0.001*

1500

Phenanthrene concentration (mg/kg soil) Fig. 2. Mean (n ¼ 10) % change in bucket intensities of A) isoleucine, B) alanine, C) glutamine and D) maltose in phenanthrene-exposed earthworms. The p values for statistically significant changes relative to the control intensities (from two sample ttests, comparing each exposed earthworm set to the control earthworm set) are labeled (*p < 0.05 and þp < 0.10). Two earthworms exposed to 1000 mg/kg phenanthrene-spiked soil did not survive exposure and their data was not used for quantification. The error bars represent the propagated standard error.

S.A.E. Brown et al. / Environmental Pollution 158 (2010) 2117e2123

maltose increased in L. rubellus taken from metal-contaminated sites, but decreased in L. terrestris from the same contaminated sites. It was proposed that the variation in maltose was possibly due to differing food sources (Bundy et al., 2004). However, in our study, earthworm diet was controlled. It is possible that maltose found in the earthworm extracts could be produced by gut microflora (Bundy et al., 2004). Exposure to phenanthrene may reduce the population of microflora which would result in a decrease in the earthworms' maltose concentration. This may explain the observed decrease in mean maltose levels in E. fetida with exposure to high concentrations of phenanthrene. Furthermore, if exposure to phenanthrene resulted in decreased metabolism in the earthworms, this could lead to a decrease in glycogen/starch breakdown and thus an observed decrease in free maltose. However, there must also be further investigation into the cause of the variability in maltose concentrations [as observed in other species (Bundy et al., 2004)] to confirm maltose as an indicator of phenanthrene exposure. 3.3. Phenanthrene bioavailability and earthworm responses

Predicted total phenanthrene (mg/kg soil)

To further explore the relationship between earthworm responses and total and bioavailable phenanthrene, PLS analysis was performed. Two, eight component PLS models were created, using soxhlet-extracted phenanthrene concentrations (total phenanthrene: Q2X ¼ 0.98 and Q2Y ¼ 0.67) and HPCD-extracted phenanthrene concentrations (bioavailable phenanthrene: Q2X ¼ 0.98 and Q2Y ¼ 0.66). The phenanthrene concentration predicted by the model (based on the earthworm 1H NMR spectra) and measured soil concentrations for both total (Fig. 3A) and bioavailable (Fig. 3B) 1800

A

y = 0.87x R2 = 0.65

1600 1400 1200 1000 800 600 400 200 0 0

200

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600

800

1000

1200

1400

1600

1800

Measured total phenanthrene (mg/kg soil) 1600

Predicted bioavailable phenanthrene (mg/kg soil)

significantly (p < 0.05) in earthworms exposed to 750, 1000 and 1500 mg/kg phenanthrene compared to controls (Fig. 2C). Finally, maltose decreased significantly (p < 0.05) in earthworms exposed to 750, 1000 and 1500 mg/kg phenanthrene compared to control earthworms (Fig. 2D). The concentrations of these metabolites were also quantified (in mg/gearthworm) based on their signal intensity in the 1H NMR spectra and similar trends were observed (Supplementary material, Section S9). Our results indicate that an increase in amino acids (isoleucine, alanine and glutamine) and a decrease in maltose are potential indicators of earthworm exposure to high concentrations of phenanthrene. Furthermore, an increase in alanine may also be a potential indicator of exposure to low concentrations of phenanthrene. After exposure to contaminants such as PAHs there may be loss of homeostasis in earthworms, resulting in an increase or decrease in metabolite concentrations (Gibb et al., 1997). PAHs are believed to be metabolized in earthworms through the cytochrome P450 system (Zhang et al., 2006). Exposure to PAHs has been linked to an increase in cytochrome P450 content (Zhang et al., 2006), the production of free radicals (Brown et al., 2004) and an increase in oxygenated metabolites and oxidative stress (Owen et al., 2008). In addition to these observations, an increase in alanine production and accumulation has been reported as a “universal” stress response (Ben-Izhak Monselise et al., 2003; Forcella et al., 2007). Alanine is produced in response to radical reactions and protects cells against damage during stress events (Forcella et al., 2007). Although the exact function of alanine is unknown (Forcella et al., 2007), the observed increase in alanine concentration after phenanthrene exposure could be indicative of a general stress response and/or metabolism of phenanthrene. In addition, it is possible that phenanthrene exposure results in increased protein metabolism and thus increased amino acid concentrations, as has been demonstrated to occur with exposure to pesticides (Forcella et al., 2007; McKelvie et al., 2009; Rochfort et al., 2009). Jones et al. (2008a) identified an increase in amino acid concentrations (alanine, leucine, valine, isoleucine, lysine, tyrosine and methionine) in Lumbricus rubellus after exposure to pyrene-spiked soil (40e640 mg/kg). The observed increase in free amino acid levels was likely due to the breakdown of tissue proteins such that free amino acids could be used for metabolism (Jones et al., 2008a). Our findings are consistent with this observation. The response of L. rubellus and Lumbricus terrestris earthworms taken from soils contaminated with cadmium, copper, lead and zinc (Bundy et al., 2004) differed from that of E. fetida exposure to phenanthrene. Bundy et al. (2004) reported a slight increase in histidine concentration in L. rubellus and a decrease in histidine and 1-methylhistidine in L. terrestris. Other amino acids were not identified in earthworm responses to heavy metal exposure, suggesting that there may be distinct earthworm responses to exposure to PAHs versus metals. This is consistent with the findings of Guo et al. (2009), who found that the metabolic response of L. rubellus exposure to a heavy metal salt (cadmium) could be distinguished from the response to PAH (fluoranthene) exposure. Maltose is an intermediate formed by the breakdown of glycogen and starches during digestion (Jones et al., 2008b). It has been demonstrated that earthworms (e.g. E. fetida) are capable of breaking down starch (Prat et al., 2002; Prabha et al., 2007; Ueda et al., 2008). It has been suggested that maltose may have a role other than energy metabolism in earthworms (Bundy et al., 2002; Lenz et al., 2005), although the exact role is unclear. Other researchers have reported both increased and decreased maltose concentrations in earthworms after contaminant exposure (Bundy et al., 2002, 2004). For instance, in contact tests of Eisenia veneta earthworms exposed to 4-fluoroaniline, a decrease in maltose was observed (Bundy et al., 2002). In addition, Bundy et al. (2004) found

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B

y = 0.86x R2 = 0.64

1400 1200 1000 800 600 400 200 0 0

200

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600

800

1000

1200

1400

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Measured bioavailable phenanthrene (mg/kg soil) Fig. 3. PLS predictions and model comparison of A) measured total phenanthrene soil concentrations (mgphenanthrene/kgsoil, soxhlet extracts, 8 PLS components, Q2X ¼ 0.98, and Q2Y ¼ 0.67) and B) measured bioavailable phenanthrene (mgphenanthrene/kgsoil, HPCD extracts, 8 PLS components, Q2X ¼ 0.98, and Q2Y ¼ 0.66). Two earthworms exposed to 1000 mg/kg phenanthrene did not survive exposure and their data was not used. The error bars represent the standard error, the dotted line represents a perfect prediction and the solid line is the linear regression of the average measured versus predicted phenanthrene concentrations. The R2 values indicated on the plots refer to the linear regression of the measured versus predicted phenanthrene concentrations.

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phenanthrene fit the data well (Q2Y values of 0.67 and 0.66 respectively). External validation was also used to test the predictive ability of the PLS models. Two additional PLS models were created using w80% of the 1H NMR spectra for the soxhletextracted phenanthrene concentrations (total phenanthrene) and HPCD-extracted phenanthrene concentrations (bioavailable phenanthrene) and using the remaining w20% as an independent test set (See Supplementary material, Section S10 for details). Both models predict phenanthrene concentration well; the total phenanthrene model (Fig. S7A) had a predictive R2 value of 0.67, while the bioavailable phenanthrene model (Fig. S7B) had a predictive R2 of 0.59 (See Supplementary material, Section S10 for details). The results of this study suggest that earthworm response is correlated to both total and bioavailable phenanthrene in soil. However, because the soil used in this experiment was freshly spiked, the concentration of bioavailable (HPCD-extracted) phenanthrene was high and is not representative of an aged soil. Further research using an aged spiked soil [where bioavailability is reduced (Semple et al., 2003)] is required to confirm the observed correlation between phenanthrene bioavailability and earthworm response and to further explore the suitability of HPCD as a proxy for earthworm bioavailability in contaminated soils. Nonetheless, the observed PLS results suggest that earthworm response, as measured by 1H NMR metabolomics, is positively and quantitatively correlated to both bioavailable and total phenanthrene concentrations in soil. If these results are validated on aged contaminated soils, 1H NMR metabolomics may be suitable for measuring bioavailability of contaminated soils when prior knowledge of the concentration and type of contaminants is known. Furthermore, metabolomic studies could also have the potential to assess sites containing unknown contaminants. Future studies will focus on validating these results with time and various soil types, however, this initial study suggests that 1H NMR earthworm metabolomics may be a valuable tool for contaminated site risk assessment. 4. Conclusion Our study demonstrates that 1H NMR-based metabolomics can discriminate responses of E. fetida exposure to sub-lethal concentrations of phenanthrene in soil after only 48 h of exposure. This short exposure time offers an advantage over other direct methods of measuring contaminant uptake, such as bioassays, which generally have exposure times of several weeks (Gevao et al., 2001; Johnson et al., 2002; Barthe and Pelletier, 2007; Bergknut et al., 2007; Jonker et al., 2007; Kreitinger et al., 2007). It should be noted, however, that an unaged, spiked soil was used in this study and our study represents an initial proof of concept study that requires additional research for validation. For example, because the non-bioavailable contaminant fraction is observed to increase with time (Semple et al., 2003), the results of this study may not be representative of earthworm responses after exposure to an aged contaminated soil. However, this first study highlights the potential for 1H NMR metabolomics to measure and monitor earthworm responses and contaminant bioavailability quickly and directly. Additional research is required to test these results in aged soils and with contaminants of varying chemical and physical properties. Nonetheless, 1H NMR metabolomics holds great promise for measuring bioavailability and risks associated with exposure to sub-lethal levels of contaminants in soil. Acknowledgments We acknowledge the Province of Ontario and the Natural Sciences and Engineering Research Council (NSERC) of Canada for

supporting this research with a Premier's Research Excellence Award (MJS) and a Strategic Grant respectively. SAEB, JRM and MJS thank NSERC for a PGS-D, a PDF and a University Faculty Award (UFA), respectively. AJS and JRM thank the Government of Ontario for an Early Researcher Award (ERA) and an Ontario PDF award, respectively. We also thank Dr. Melissa Whitfield Åslund and Fiona Wong from the University of Toronto and Dr. Joshua Hicks, Dr. Klaus-Peter Neidig and Dr. Hartmut Schaefer from Bruker BioSpin for technical assistance and valuable discussions. Finally, we would like to thank the anonymous reviewers whose input greatly improved this manuscript. Appendix. Supplementary material Supplementary data associated with this article can be found in the online version at doi:10.1016/j.envpol.2010.02.023. References Baroni, M., Costantino, G., Cruciani, G., Riganelli, D., Valigi, R., Clementi, S., 1993. Generating optimal linear PLS estimations (GOLPE): an advanced chemometric tool for handling 3D-QSAR problems. Quantitative Structure-Activity Relationships 12, 9e20. Barthe, M., Pelletier, E., 2007. Comparing bulk extraction methods for chemically available polycyclic aromatic hydrocarbons with bioaccumulation in worms. Environmental Chemistry 4, 271e283. Ben-Izhak Monselise, E., Parola, A.H., Kost, D., 2003. Low-frequency electromagnetic fields induce a stress effect upon higher plants, as evident by the universal stress signal, alanine. Biochemistry and Biophysical Research Communications 302, 427e434. Bergknut, M., Sehlin, E., Lundstedt, S., Andersson, P., Haglund, P., Tysklind, M., 2007. Comparison of techniques for estimating PAH bioavailability: uptake in Eisenia fetida, passive samplers and leaching using various solvents and additives. Environmental Pollution 145, 154e160. Bonin, J.L., Simpson, M.J., 2007. Variation in phenanthrene sorption coefficients with soil organic matter fractionation: the result of structure or conformation? Environmental Science and Technology 41, 153e159. Brinch, U., Ekelund, F., Jacobsen, C.S., 2002. Method for spiking soil samples with organic compounds. Appl. Environ. Microbiol. 68, 1808e1816. Brown, P.J., Long, S.M., Spurgeon, D.J., Svendsen, C., Hankard, P.K., 2004. Toxicological and biochemical responses of the earthworm Lumbricus rubellus to pyrene, a non-carcinogenic polycyclic aromatic hydrocarbon. Chemosphere 57, 1675e1681. Brown, S.A.E., Simpson, A.J., Simpson, M.J., 2008. Evaluation of sample preparation methods for 1H NMR metabolic profiling studies with Eisenia fetida. Environmental Science and Technology 27, 828e836. Bundy, J.G., Keun, H.C., Sidhu, J.K., Spurgeon, D.J., Svendsen, C., Kille, P., Morgan, A.J., 2007. Metabolic profile biomarkers of metal contamination in a sentinel terrestrial species are applicable across multiple sites. Environmental Science and Technology 41, 4458e4464. Bundy, J.G., Lenz, E.M., Bailey, N.J., Gavaghan, C.L., Svendsen, C., Spurgeon, D.J., Hankard, P.K., Osborn, D., Weeks, J.M., Traugere, S.A., Speir, P., Sanders, I., Lindon, J.C., Nicholson, J.K., Tang, H., 2002. Metabonomic assessment of toxicity of 4-fluoroaniline, 3, 5-difluoroaniline and 2-fluoro-4-methylaniline to the earthworm Eisenia veneta (Rosa): identification of new endogenous biomarkers. Environmental Toxicology and Chemistry 21, 1966e1972. Bundy, J.G., Osborn, D., Weeks, J.M., Lindon, J.C., Nicholson, J.K., 2001. An NMR-based metabonomic approach to the investigation of coelomic fluid biochemistry in earthworms under toxic stress. FEBS Letters 500, 31e35. Bundy, J.G., Spurgeon, D.J., Svendsen, C., Hankard, P.K., Weeks, J.M., Osborn, D., Lindon, J.C., Nicholson, J.K., 2004. Environmental metabonomics: applying contamination biomarker analysis in earthworms at a metal contaminated site. Ecotoxicology 13, 797e806. Cramer III, R.D., 1993. Partial Least Squares (PLS): its strengths and limitations. Perspectives in Drug Discovery and Design 1, 269e278. Dean, J.R., Scott, W.C., 2004. Recent developments in assessing the bioavailability of persistent organic pollutants in the environment. Trends in Analytical Chemistry 23, 609e618. Doick, K.J., Clasper, P.J., Urmann, K., Semple, K.T., 2006. Further validation of the HPCD-technique for the evaluation of PAH microbial availability in soil. Environmental Pollution 144, 345e354. Dunn, W.B., Ellis, D.I., 2005. Metabolomics: current analytical platforms and methodologies. Trends in Analytical Chemistry 24, 285e294. Eijsackers, H., van Gestel, C.A.M., de Jonge, S., Muijs, B., Slijkerman, D., 2001. Polycyclic aromatic hydrocarbon-polluted dredged peat sediments and earthworms: a mutual interference. Ecotoxicology 10, 35e50. Ekman, D.R., Teng, Q., Villeneuve, D.L., Kahl, M.D., Jensen, K.M., Durhan, E.J., Ankley, G.T., Collette, T.W., 2008. Investigating compensation and recovery of

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