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Soil Biology & Biochemistry 78 (2014) 21e29

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Soil Biology & Biochemistry

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The influence of earthworm and mycorrhizal co-inoculation on Cdspeciation in a contaminated soil

Fatemeh Aghababaei*, Fayez Raiesi, Alireza HosseinpurDepartment of Soil Sciences and Engineering, Faculty of Agriculture, Shahrekord University, P.O. Box 115, Shahrekord, Iran

a r t i c l e i n f o

Article history:Received 28 December 2013Received in revised form31 May 2014Accepted 13 June 2014Available online 3 July 2014

Keywords:Earthworm activityMycorrhizal symbiosisCd fractionationCd bioaccumulation

* Corresponding author.E-mail addresses: aghababaei_fateme90@yahoo.c

com (F. Aghababaei).

http://dx.doi.org/10.1016/j.soilbio.2014.06.0100038-0717/© 2014 Elsevier Ltd. All rights reserved.

a b s t r a c t

This experiment was conducted to establish how earthworms and arbuscular mycorrhizal fungi (AMF)interactively impact cadmium (Cd) chemical fractions in a calcareous soil artificially contaminated withCd. The chemical forms of Cd using the Sposito's sequential extraction procedure were determined in asoil spiked with Cd (10 and 20 mg Cd kg�1), inoculated or un-inoculated with earthworm (Lumbricusrubellus L. species) and AMF (Glomus intraradices and Glomus mosseae species) under greenhouse con-ditions for two months. Results showed that Cd concentrations of the non-residual fraction (i.e., soluble/exchangeable, organic bound and inorganic bound forms) increased with earthworm addition and theresidual Cd fraction tended to decrease. Arbuscular mycorrhizal inoculation decreased the inorganicbound Cd fraction with a concurrent increase in the residual Cd fraction. However, no significant in-teractions between earthworm and AMF were observed for the non-residual Cd fraction, suggesting thecombined influence of both soil organisms on the easily and potentially available Cd was largely inde-pendent or additive. While the presence of both earthworms and AMF resulted in an antagonisticinteraction on the residual Cd fraction at high Cd level, this interaction was additive at low Cd level.Cadmium addition increased its uptake into the body of earthworms, which was significantly greater athigh than low Cd levels. However, bioaccumulation factor for Cd accumulation in the earthworm bodywas lower at high than low Cd levels, indicating the low transfer of Cd from the soil environment toearthworm tissues at higher exposure levels. It is concluded that earthworms affects the chemical Cdfractions more than AMF.

© 2014 Elsevier Ltd. All rights reserved.

1. Introduction

Cadmium (Cd), a potentially toxic metal, enters the food chainthrough plant uptake from polluted soils (Kirkham, 2006; Smith,2009).Undoubtedly, a high level of this highly mobile and toxicmetal in the soil is a major danger to both the environmentalquality and human health in the long-term (Ali et al., 2013). Cad-mium is well-known to have an adverse impact on all the livingorganisms in soil and terrestrial plants in Cd-polluted environ-ments. However, the potential toxicity of Cd for soil macro- andmicro-biota, and crop growth often is more closely correlated withthe concentration of bio-available Cd fraction than with the totalmetal concentration (Vig et al., 2003; Kirkham, 2006). Thus, it isindispensable to understand changes in Cd mobility and bioavail-ability, which depends largely on its chemical forms in the soil

om, faghababaei1980@gmail.

rather than on the total Cd level (Sposito et al., 1982; Ma and Rao,1997).

The sequential extraction or solid phase speciation techniquesfor Cd fractionation offer a powerful means for assessing its dis-tribution with different bioavailability in soil (Sposito et al., 1982;Ma and Rao, 1997). Cadmium fractionation usually results inoperationally defined fractions or forms that are closely linked to itsbioavailability to soil organisms and plants (Nannoni et al., 2011;Bai et al., 2008). Several abiotic and biotic factors such as soil pH,soil organic matter (SOM) and dissolved organic carbon (DOC)contents, cation exchange capacity, fertilizer application and soilorganisms directly or indirectly affect Cd speciation and conse-quently its mobility and bioavailability (Adriano, 2001; Vig et al.,2003; Bolan et al., 2003; Renella et al., 2004; Kirkham, 2006;Pelfrene et al., 2012).

Among soil organisms, arbuscular mycorrhizal fungi (AMF) andearthworms are the reported biotic factors which may have aninfluence on the mobility and bioavailability of Cd and other heavymetals in soileplant systems, largely owing to changes in their

F. Aghababaei et al. / Soil Biology & Biochemistry 78 (2014) 21e2922

distribution among various soil fractions (Tao et al., 2003; Wenet al., 2004; Dai et al., 2004; Lukkari et al., 2006; Huang et al.,2005, 2008; Bai et al., 2008; Ruiz et al., 2011). Specifically, bothearthworms (Wen et al., 2004; Ruiz et al., 2009; Nannoni et al.,2011) and mycorrhizal fungi (Giasson et al., 2005; Huang et al.,2008) have been found to influence soil Cd behavior and chemis-try with changes in its immobilization and mobilization or itsspeciation in Cd-polluted soils. However, the individual effects ofearthworms and AMF on the availability and speciation of Cd andother toxic metals are inconsistent in the literature when earth-worms (Wen et al., 2004; Sizmur et al., 2011; Ruiz et al., 2009;Natal-da-Luz et al., 2011; Kizilkaya, 2004; Ruiz et al., 2011) orAMF (Munier-Lamy et al., 2007; Bai et al., 2008; Huang et al., 2008;Subramanian et al., 2009) were present.

Earthworms may often modify individually the availability andspeciation of Cd and other toxic metals in soil by changing the soilcharacteristics such as pH and DOC, and the stimulation of micro-bial activity (Ma et al., 2002; Wen et al., 2004; Sizmur et al., 2011).Ma et al. (2002) reported that available Pb and Zn concentrationsincreased in earthworm-worked soils. The addition of earthwormsaffected Zn, Pb, As and Cu fractionations, to different extents,depending on soil type and heavy metals (Cheng and Wong, 2002;Kizilkaya, 2004; Sizmur et al., 2011). An increase of Cd and othermetals (Cr, Co, Ni, Zn, Cu, and Pb) in thewater soluble, exchangeableand carbonate fractions was reported when Eisenia fetida waspresent (Wen et al., 2004). The activity of E. fetida in soils pollutedby mining activities changed the chemical forms of Cd and othermetals (Pb, Zn, Cu) with a significant increase in the non-residualfractions (Ruiz et al., 2009). On the other hand, the activity ofearthworms (Pheretima sp.) declined the concentration ofexchangeable and carbonate Zn fractions, however, the significanceof the changes depended upon the soil type (Ma et al., 2002). Theinfluence of Lumbricus terrestris on Pb and Zn speciationwas minor,with a significant increase only in organically bound heavy metalsin mining soils (Ruiz et al., 2011).

AMF may also influence Cd mobility and toxicity by increasingsoil pH (Shen et al., 2006), sequestering Cd inside extra-radicalmycelium (Janou�skov�a and Pavlíkov�a, 2010) and binding Cd ionsto glomalin (Gonz�alez-Ch�avez et al., 2004; Aghababaei et al., 2014),a glycoprotein produced by AMF. Arbuscular mycorrhizal hyphaecan change Cd from a carbonate to a water-soluble form, sincefungal hyphae exude simple organic acids like citric and oxalic acidsthat facilitate the solubility of heavy metals (Giasson et al., 2005).AMF inoculation resulted in the transformation of unavailable Znforms into available forms by increasing the organically bound Znfraction and reducing crystalline oxide and residual Zn fractions(Subramanian et al., 2009). The speciation of heavy metals (Cu, Znand Pb) was changed by AMF symbiosis from bio-available to non-bio-available forms (Huang et al., 2005). In contrast, mycorrhizalinoculation by Glomus mosseae species did not affect Se speciation(Munier-Lamy et al., 2007). All these discrepancies in the publishedresults could stem from differences in earthworm and fungal spe-cies, pollution mode, heavy metals and soil conditions, and there-fore cannot be an overall impact of earthworm and AMFinoculation.

Although the overall influences of earthworms and AMF aloneon the speciation of Cd and other metals have been studied incontaminated soils with variable results, to our knowledge no studywas conducted to show their interactive effects. The presence ofearthworm activity and AMF has been proposed as a practice forboth bioremediation of polluted soils and plant protection againstmetal toxicity, but their combined effects on toxic metal speciationare not clear. The influence of these organisms on metal speciationmight be particularly crucial in contaminated soils, since they mayconcurrently change heavy metal availability and mobility with a

consequence for their uptake by plant and phyto-extraction incontaminated environments (Yu et al., 2005; Ma et al., 2006; Huaet al., 2010). However, all these studies did not attempt to deter-mine changes in the metal speciation with earthworm activity andAMF inoculation. Specifically, we should know how the biotic in-teractions between these two soil organisms affect Cd availabilityand toxicity, and whether the combined effects are additive orinteractive. This can further enhance our understanding of the in-teractions between earthworms and AMF and their significant rolein either protecting plants against the phyto-toxicity of metals or inpromoting phyto-remediation efficiency in soils polluted with toxicmetals. Therefore, the main aim of the current study was toinvestigate changes in Cd speciation in a calcareous soil artificiallyspiked with 10 and 20 mg Cd kg�1 and inoculated with earthworm(Lumbricus rubellus) and AMF (Glomus intraradices and G. mosseaespecies) under glasshouse conditions. We hypothesized thatearthworms and AMF would interactively affect Cd speciation andtransformation. It is assumed that earthworm activity wouldmodify AMF effects on soil Cd fractionation pattern or AMF inoc-ulation would change earthworm effects on the distribution of Cdamong soil solid phases.

2. Materials and methods

2.1. Experimental lay-out

The experiment was 3 � 2 � 3 factorial with Cd level, earth-worm and AMF inoculation arranged in a completely randomizeddesign with three replications. Treatments consisted of a fullfactorial combinations of three Cd levels (0, 10 and 20 mg kg�1)applied as CdCl2, two earthworm treatments (no earthworm, NE,and with L. rubellus earthworm, WE) and three AMF treatments (G.intraradices, G. mosseae, and non-mycorrhizal control, NM) in acalcareous soil cropped with maize (Zea mays L.) under glasshouseconditions for two months.

2.2. Soil preparation and treatment

A typical calcareous sandy loam soil, classified as Typic Calcix-erepts (Soil Survey Staff, 2010), from the 0e30 cm layer was ob-tained in a cropland fieldwithout pollution history. The soil was air-dried, passed through a 2mm sieve and autoclaved at 121 �C for 2 h.A subsample of the study soil was analyzed for general chemicaland physical properties. The soil had the following physical andchemical properties: pH (in H2O) 8.1, ECe 0.20 dS m�1; CEC19.7 cmol (þ) kg�1; CaCO3 190 mg g�1; organic C 2.8 mg g�1, total N0.6 mg g�1, available P and K 5.7 and 168mg kg�1, respectively; clay16%, silt 12% and sand 72%. We used plastic pots with the bottomcovered with nylon net to prevent earthworm escape. In total, 54pots were prepared and filled with 8 kg (fresh weight) autoclavedsoil. To establish and reactivate soil microorganisms, 50 ml soilsuspension was added to each pot to inoculate the autoclaved soilwith fresh microorganisms. To obtain the soil suspension, 500 gfresh soil was suspended in 1.5 L de-ionized water and filteredthrough a 25 mm for eliminating AMF spores (Schroeder and Janos,2004). The soils were artificially contaminated with Cd (as cad-mium chloride) at the following rates: 0 (control), 10 and20 mg kg�1 on a dry weight basis. The control treatments werewatered with distilled water and others with appropriate aliquotsof aqueous solutions of Cd chloride to obtain the above concen-trations. Soils were mixed thoroughly for an even distribution ofadded Cd in the soil matrix. De-ionized distilled water was added tothe soil to achieve a moisture content of about 60e70% of fieldcapacity. The soils were incubated at room temperature (about20 �C) for 4 months, allowing Cd to distribute into various fractions,

F. Aghababaei et al. / Soil Biology & Biochemistry 78 (2014) 21e29 23

to reach an equilibrium condition, and to reduce the effects ofsampling disturbance. During the incubation period, the pots wereweighed periodically to monitor soil water losses, and moisturelosses were replaced by tap water.

2.3. Earthworm and AMF treatments

Epigeic earthworms (L. rubellus L. Savigny) were collected byhand from a garden soil, transferred to the laboratory and allowedreproduction for four months in plastic boxes filled with soil andplant residues. The AMF inoculums (G. intraradices and G. mosseae)consisted of mixed rhizosphere soil samples from maize culturescontaining 50 spores of each fungus per gram, hyphae and heavilyinfected root fragments with many fungal spores. Maize (Zea maysL.) seeds were surface-sterilized for 5 min in 5% NaClO and washedin sterilized distilledwater andwere germinated onwet filter paperin petri dishes. After germination, the seedlings were selected foruniformity and transplanted into pots. At transplant, maize seed-lings were inoculated with 3 g of AMF inoculums, which was placedunder the roots. At the end of soil incubation, two pre-germinatedseedlings were sown in each pot. About 60 g of each AMF inoculum(G. intraradices and G. mosseae) was added to one-third of the potsassigned to each mycorrhizal treatment. Earthworms were placedin Petri dishes and kept in sterilized glass vessels for 24 h in tapwater to empty their guts. They were washed with distilled waterto remove surface soils and minimize the number of naturallyoccurring mycorrhizal propagules associated with their surfaces orgut contents. Four adult earthworms (with a well-developedclitellum) with similar fresh weights (0.4 ± 0.06 g) and lengths(6 ± 0.80 cm) were added to half of the pots 2 weeks after maizecultivation. Alfalfa residue was added at a rate of 15 g per pot asfeed for earthworms.

2.4. Soil analysis

After plant harvest, soil sub-samples were taken for Cd frac-tionation. Earthworms were collected manually, washed carefullyand their gut contents were emptied for two days, oven-dried at105 �C for one day and weighed for measuring dry biomass. Thedried earthworms were digested overnight in nitric acid (1 mlHNO3 per mg dry weight). After heating at 120 �C, 1 ml HNO3/H2SO4/HCl (10/2/3; v/v/v) was added. The solution was heated at180 �C; and was finally diluted with deionized water up to 25 mlafter cooling. The concentration of Cd in earthworm extracts wasdetermined by an atomic absorption spectrophotometer (AASModel GBC 913 plus) with a detection limit of 0.0009 mg L�1. Thesequential extraction method outlined by Sposito et al. (1982) wasused to fractionate the added Cd into Cd pools of different mobilityand bioavailability. This approach allows the separation of soil Cdby a four-step sequential extraction procedure into soluble andexchangeable Cd (0.5 M KNO3, 16 h), organically bound Cd (0.5 mNaOH, 16 h), Cd in inorganic precipitates (i.e., oxides or carbonates)or inorganically bound Cd (0.05 M Na2EDTA, 6 h) and residual orsulfide Cd (4 M HNO3) fractions (Sposito et al., 1982; Bolan et al.,2003; Liu et al., 2009). The first three Cd species were consideredas the non-residual fractionwith different mobility. The soluble andexchangeable Cd fractions are thought to be the most mobile andimmediately plant-available forms (Amir et al., 2005; Jalali andArfania, 2011), while organically and inorganically bound Cd frac-tions can be considered relatively mobile, depending on environ-mental conditions (Amir et al., 2005). The residual (HNO3-extracted) Cd fraction is assumed to be in insoluble sulfideminerals(Sposito et al., 1982) but mobilizable (Amir et al., 2005) or incor-porated into the crystalline lattice of soil minerals, which can bequite immobile (Jalali and Arfania, 2011). The relative distribution

of Cd fractions was calculated as the concentration of each Cdfraction divided by the sum of the four Cd fractions extracted withthis sequential extraction scheme.

Cadmium concentration in soil extracts was measured usingAAS. The results of Cd contents and distributions were expressed onan oven-dry weight basis. We calculated bioaccumulation factor(BAF) for the accumulation of Cd in earthworm body to quantify thebioavailability of Cd as:

BAF ¼ CdðearthwormÞCdðsoilÞ

where Cd(earthworm) is the concentration of Cd in the earthwormtissues and Cd(soil) is the concentration of Cd added to the soil(Nannoni et al., 2011; Van Gestel et al., 2011).

2.5. Statistical analyses

Soil data were subjected to factorial analysis of variance(ANOVA). Prior to ANOVA procedure, data were checked for normaldistribution and equal variance. The effects of independent factors(Cd level, earthworm addition, and AMF inoculation) on soil Cdfractions were analyzed using 3 � 2 � 3 full factorial three-wayANOVA. F-values and significance levels for three-way ANOVA ofthe data were analyzed using General Linear Model (GLM) inMinitab16 software. When statistically significant, the mean valuesfor the main and interactive effects of the factors were separated bythe post-hoc Tukey HSD test (P � 0.05).

3. Results and discussion

3.1. Cd, earthworm and AMF main effects on soil Cd fractions

Overall, the average recovery of the added Cd was calculated bycomparing the sum of the four Cd fractions with the total Cd added.Cadmium recovery ranged from 96 to 97% at the two Cd additionrates, which is a good percentage recovery for the Cd initially addedto the study soil. The low Cd recovery is most probably due to Cdleached out of the soils and uptake intomaize plant and earthwormtissues, which were not considered in the recovery calculation.ANOVA results for Cd concentration (mg kg�1) and distribution (%)in different soil components as affected by the addition of Cd,earthworm and AMF are presented in Table 1. In general, a signif-icant proportion (54e57%) of the added Cd was present in theinorganic precipitates (i.e., carbonates and Fe/Mn oxides) in thiscalcareous soil (Table 2). As a result, it is possible that Cd mobility isreduced by its adsorption on soil carbonates when soil is artificiallyspiked with this metal. The residual fraction was the next mostimportant pool (35e42%) for the added Cd. Similar findings werereported by Renella et al. (2004) andMoral et al. (2005), who foundthe main proportion of the added Cd to calcareous soils wasmeasured in carbonates and residual fractions. Addition of Cd tothis calcareous soil resulted in its accumulation in carbonate-boundfraction, probably due to the high pH value (8.1) and carbonatecontent (19%). High soil pH and carbonates decline Cd solubility,and consequently its mobility (Renella et al., 2004; Strobel et al.,2005). Only a minor proportion of the added Cd was in solubleand exchangeable fraction (2.8e5.7%) at the end of the experimentand organic bound Cd was the smallest fraction (0.6e1.4%) in thissoil (Table 2), which is consistent with other observations (Renellaet al., 2004; Moral et al., 2005). The Cd concentration of organicfraction in this soil was relatively low, due largely to low organic Ccontent (0.28%). This data indicates that the easily extractable andexchangeable Cd fractions together represent less than 6%, which is

Table 1The 3 � 2 � 3 factorial ANOVA results (F-values) showing the influence of Cd level, earthworm (EW) and arbuscular mycorrhizal fungi (AMF) on Cd chemical fractions in soiland on earthworm properties.

Variable Main effects Interaction effects C.V.

Cd EW AMF Cd � EW Cd � AMF EW � AMF Cd � EW � AMF

df 2 1 2 2 4 2 4 e

Soluble/Exchangeable Cd 293*** 20.6*** ns 6.44** ns ns ns 8.09Organic Cd 1116*** 111*** ns 62.1*** ns ns ns 20.4Inorganic Cd 3925*** 12.8** 13.1** 24.4*** 3.45* 3.43* ns 10.3Residual Cd 9179*** 56.7*** 5.70** 57.8*** 8.96** ns 11.7** 10.5Earthworm dry weight 12.8** 823*** ns 12.8** ns ns ns 31.3Earthworm Cd uptake 19589*** 51464*** ns 19589*** ns ns ns 92.6Earthworm Cd BAF 6550*** 24948*** ns 6550*** ns ns ns 75.5

*P < 0.05; **P < 0.01; ***P < 0.001, ns, not significant.BAF ¼ bioaccumulation factor.

Table 2Effects of Cd level and earthworm (NE ¼ no earthworm, WE ¼ with earthworm) onCd chemical fractions in contaminated soils cropped with maize under greenhouseconditions.

Cd added(mg kg�1)

EW treatment Mean EW treatment Mean

NE WE NE WE

Soluble/ExchangeableCd (mg Cd kg�1 soil)

Soluble/Exchangeable Cd (%)

0 ND aB ND aB ND B ND aC ND aC ND C10 0.517bA 0.584aA 0.551A 5.35bA 6.03aA 5.69A20 0.531aA 0.568aA 0.549A 2.75aB 2.89aB 2.82BMean 0.524b 0.576a 4.05b 4.46a

Organic Cd (mg Cd kg¡1 soil) Organic Cd (%)0 ND aC ND aC ND B ND aC ND aC ND C10 0.097bB 0.163aA 0.130A 1.03bA 1.68aA 1.35A20 0.117aA 0.131aB 0.124A 0.604aB 0.671aB 0.64BMean 0.107b 0.147a 0.817b 1.18a

Inorganic Cd (mg Cd kg�1 soil) Inorganic Cd (%)0 ND aC ND aC ND C ND aC ND aB ND C10 5.69aB 5.44aB 5.56B 58.5aA 56.0aA 57.3A20 9.90bA 11.2aA 10.5A 51.2bB 57.1aA 54.2BMean 7.80b 8.32a 54.8b 56.5a

Residual Cd (mg Cd kg�1 soil) Residual Cd (%)0 ND aC ND aC ND C ND aC ND aC ND C10 3.37aB 3.35aB 3.36B 35.4aB 34.6aB 35.0B20 8.72aA 7.60bA 8.16A 45.1aA 38.7bA 41.9AMean 6.05a 5.47b 39.8a 36.6b

Within each row, mean values with the same lowercase letter are not significantlydifferent (P > 0.05) between EW treatments by the Tukey test (n ¼ 27 for EWtreatment, n ¼ 9 for Cd � EW).Within each column, mean values with the same uppercase letter are not signifi-cantly different (P > 0.05) among Cd levels by the Tukey test (n ¼ 18 for Cd treat-ment, n ¼ 9 for Cd � EW).ND, not detectable by atomic absorption spectrophotometry (detection limit:0.0009 mg L�1 equivalent to 0.02 mg kg�1 soil after extraction).EW ¼ earthworm.

F. Aghababaei et al. / Soil Biology & Biochemistry 78 (2014) 21e2924

not attractive for Cd phyto-remediation processes such as phyto-extraction and phyto-stabilization. However, a major proportion(>55%) of the added Cd was associated with other non-residualfractions (i.e., inorganically and organically bound). This wouldmean the possibility for the potential mobility and bioavailability ofCd in the study soil as a consequence of changes in environmentalfactors enhancing Cd solubility and availability.

3.1.1. Cd effectCadmium addition significantly increased the concentration of

all Cd fractions across earthworm and AMF treatments (Tables 2and 3). The increase in soluble/exchangeable and organic boundCd fractions was equal in both Cd-polluted soils, but greater at highthan low Cd levels for both inorganic bound and residual Cd frac-tions. This indicates redistribution of the added Cd to various soil

solid phases, which is consistent with the observation reported byBolan et al. (2003) in variable change soils. The relative distributionof Cd among the chemical forms also depended on the added levelof this element (Table 2). The relative proportions of Cd in thesoluble/exchangeable, organic bound and inorganic bound Cdfractions were greater at low than high Cd additions. As with Cdconcentration, the relative contribution of the residual fraction tothe total Cd was higher at high than low Cd additions. Thesecorrespond with the results reported by others (Bolan et al., 2003;Renella et al., 2004).

3.1.2. Earthworm effectWe found that the main effect of earthworm treatment on soil

Cd fractionation pattern was significant (Table 1). Earthwormaddition increased the concentrations of soluble/exchangeable,organic bound and inorganic bound Cd fractions by 10, 37 and 7%,respectively; with a concomitant decline of the residual (sulfide)fraction (9.6%) when compared with the control soil withoutearthworm (Table 2). Likewise, the relative distribution of Cd spe-cies in earthworm-worked soils was significantly different fromthat in earthworm not-worked soils, and showed a trend almostsimilar to the concentration values. Our results demonstrate thepotential capability of earthworms to enhance the mobility andbioavailability of the added Cd primarily by an increase in thesoluble/exchangeable fraction. Several reasons have been attrib-uted to the earthworm-induced enhancement of heavy metalmobility and availability (Sizmur and Hodson, 2009) and increasesof less bioavailable fractions due to earthworm activity (Cheng andWong, 2002; Wen et al., 2004; Lukkari et al., 2006; Udovic andLestan, 2007, 2010; Ruiz et al., 2009; Sizmur et al., 2011; Ruizet al., 2011); decreases in soil pH; increases in DOC concentrationand enhanced microbial activity in the earthworm gut and casts.Although the soil pH increased by 0.15 units from 7.74 in pots withearthworm activity to 7.89 in pots without earthworm (data notshown), concentration of the soluble/exchangeable Cd fraction wasenhanced with earthworm addition. This indicates that soil pH isvery unlikely to affect the soluble and exchangeable Cd fractionswith earthworm activity in this study. However, the impact ofearthworms on soil pH and metal availability is often variable,partly due to differences in earthworm species and soil conditions(Wen et al., 2004). For example, the soil pH remained unaffected byL. rubellus activity and increased by epigeic E. fetida in EDTA-untreated contaminated soils, while increasing with the activityof both earthworm species in EDTA-amended soils without a sig-nificant change in Pb and Zn fractionation (Udovic and Lestan,2007). According to Wen et al. (2004), earthworm (E. fetida) inoc-ulation resulted in an increase in soil pH with a concurrent raise inthe concentration of water-soluble heavy metals (Zn, Cu, Cr, Cd, Co,Ni, and Pb). Therefore, it appears that there must be other

Table 3The concentration and distribution of residual Cd fraction in earthworm and AMF(Glomus intraradices and Glomus mosseae species) inoculated soils with increasingCd concentration.

Treatment Cd added (mg kg�1) Mean (EW) Mean (AMF)

EW AMF 0 10 20

mg Cd kg�1 soilNE NM ND cA 3.62bA 7.98aB 6.05A 5.58B

G. intraradices ND cA 3.22bA 9.06aA 5.89AG. mosseae ND cA 3.29bA 9.12aA 5.81A

WE NM ND cA 3.20bA 7.54aB 5.47BG. intraradices ND cA 3.43bA 7.84aBG. mosseae ND cA 3.42bA 7.41aB

Mean (Cd) ND c 3.35b 8.16a%NE NM ND cA 34.8bA 44.6aAB 39.8A 38.0A

G. intraradices ND cA 35.4bA 44.9aA 38.7AG. mosseae ND cA 34.8bA 44.6aA 38.4A

WE NM ND cA 34.3bA 38.3aB 36.6BG. intraradices ND bA 35.2aA 39.4aBG. mosseae ND bA 34.9aA 39.3aB

Mean (Cd) ND c 35.0b 41.9a

Within each row, mean values (n ¼ 3) with the same lowercase letter are notsignificantly different (P > 0.05) among Cd levels by the Tukey test.Within each column, mean values (n ¼ 3) with the same uppercase letter are notsignificantly different (P > 0.05) among AMF treatments by the Tukey test.ND, not detectable by atomic absorption spectrophotometry (detection limit:0.0009 mg L�1 equivalent to 0.02 mg kg�1 soil after extraction).EW ¼ earthworm treatment, AMF ¼ arbuscular mycorrhizal fungi, NE ¼ no earth-worm, WE ¼ with earthworm, NM ¼ non-mycorrhizal.

F. Aghababaei et al. / Soil Biology & Biochemistry 78 (2014) 21e29 25

mechanisms for the enhanced Cd mobility and availability inearthworm-worked soils. Changes in soil DOC or soluble organiccarbon have been found to regulate the solubility and availability ofheavy metals in polluted soils (Antoniadis and Alloway, 2002; Wenet al., 2004; Wong et al., 2007; Sizmur and Hodson, 2009). Theincrease in DOC can instigate desorption of heavy metals byaffecting the heavy metal adsorption-desorption equilibria andmetal binding to a variety of low molecular weight organic com-ponents in this soluble carbon pool (Antoniadis and Alloway, 2002;Wong et al., 2007). Natural DOC derived from sewage sludgeincreased Cd mobility through the formation of soluble DOC-Cdcomplexes in two different Brown Earth soils (Antoniadis andAlloway, 2002) and calcareous soils (Wong et al., 2007). Earth-worm activity can enhance soil DOC contents due to their effects onmicrobial activity and SOM mineralization (Wen et al., 2004;Sizmur and Hodson, 2009; Sizmur et al., 2011). Our previousexperiment (Aghababaei et al., 2014) also indicated a significantincrease in DOC concentration with anecic L. rubellus activity incalcareous soils polluted by Cd. Likewise, the addition of E. fetida tothe soil increased water-soluble C and water-soluble carbohydrateconcentrations of a sandy soil (Caravaca and Rold�an, 2003). It ap-pears that increased DOC content is likely to be the reason for theenhanced soluble and exchangeable Cd fractions with earthworminoculation in the study soil. Similar findings were reported byWenet al. (2004) and Aghababaei et al. (2014), who showed theincreased DOC by earthworm activity may contribute to an increasein the concentration of Cd and other heavymetals (Cr, Co, Ni, Zn, Cu,and Pb) in the water soluble/exchangeable fraction. The highersolubility of Cu, Zn and Pb in earthworm casts compared with bulksoil was attributed to the greater concentration of soluble organiccarbon in the earthworm casts compared to bulk soil (Sizmur et al.,2011). A significant correlation between DOC contents and theconcentration of water extractable heavy metals (Wen et al., 2004)and rare earth elements (Wen et al., 2006) was observed in thepresence of E. fetida in Chinese soils, which supports our results. Incontrast, addition of the burrowing earthworm (Pheretima

guillelmi) to mine tailings declined or had no impact on the mobilefraction of metals such as Pb and Zn, largely depending upon thelevel of metals in tailings and the extracting solution used (Ma et al.,2006). However, to determine a relationship between the increasedplant-available Cd as a result of earthworm activity and its subse-quent uptake into plants, bioassay tests reflecting earthworm im-pacts on the soluble/exchangeable fraction are required.

Our results indicate that earthworm activity increased bothorganic and inorganic bound Cd fractions, but the increase in theorganic Cd forms (37%) was greater than that in the inorganic forms(7e10%). Similarly, organic-Zn fraction was significantly increasedby the activity of L. rubellus species in remediated soils (Udovic andLestan, 2007) and Pheretima sp. in paddy soils (Cheng and Wong,2002) both using the Tessier's sequential extraction procedure.Inoculation of L. terrestris resulted in a significant increase inorganically bound Pb and Zn in mining soils using the BCR (Com-munity Bureau of Reference) sequential extraction method (Ruizet al., 2011). The increased organic-Cd fraction is probably due tothe stimulation of SOM mineralization and humification by earth-worm burrowing, feeding and casting activities (Edwards andBohlen, 1996). Earthworms can increase soil microbial populationand metabolic activity by producing organic materials (Binet et al.,1998;Wen et al., 2004) with a consequence for SOM decompositioninto smaller and low molecular weight water-soluble components.Earthworms could also increase the organic-metal fraction byexcreting organic compounds and humified organic matter (Brownet al., 2000), which may form strong complexes with metals(Kizilkaya, 2004). In the current study, Cd was predominantlypresent in the organic forms of the non-residual fraction, which canbe potentially bioavailable since the organic components are easilydecomposed by soil microorganisms with the subsequent increasein the concentration of metal ions into soil solution (Cheng andWong, 2002). The Cd associated with organic particles might alsobe released after the soil transit through the gut of earthworms, andwith enhanced microbial activity. With SOM mineralization in thealimentary canal of the earthworm (Edwards and Bohlen,1996), themetals previously bound to organic matter would indeed bereleased slowly over time.

Earthworm activity was also found to increase soil calciumcontent due to excretion of calcite into the soil by calciferous glands(Edwards and Bohlen, 1996; Brinza et al., 2014; Versteegh et al.,2014), probably increasing the carbonate bound Cd fractionthrough addition of calcium compounds to the study soil. In arecent study, Brinza et al. (2014) showed that soil Zn was incor-porated into calcium carbonate granules secreted by the earth-worm L. terrestris during their formation. Our finding is inagreement with Wen et al. (2004), who observed an increase of Cdand other metals (Cr, Co, Ni, Zn, Cu, and Pb) in the carbonate, watersoluble and exchangeable fractions sequentially extracted by thethree-step BCR method due to the activity of E. fetida. Similarly, theactivity of E. fetida in soils polluted by mining activities affected thechemical forms of Cd and other heavy metals (Pb, Zn, Cu) withsignificant increases in the non-residual fractions of the soil (Ruizet al., 2009). The higher pH in earthworm's gut due to the excre-tion of ammonia and/or calcium carbonate could also induce metalprecipitation with carbonates (Udovic and Lestan, 2010). Never-theless, in some studies, addition of earthworms has been shown toresult in no clear effect on metal fractionation patterns with de-creases or no change in metals bound to either organic matter orcarbonates (Wen et al., 2004; Ruiz et al., 2011), contrasting to ourresults. For example, the impact of L. terrestris on Pb and Znspeciationwas minor, with a significant increase only in organicallybound heavy metals in mining soils using the BCR sequentialextraction scheme (Ruiz et al., 2011). The addition of earthworms(Pheretima sp.) decreased the concentration of Zn associated with

F. Aghababaei et al. / Soil Biology & Biochemistry 78 (2014) 21e2926

carbonates, although the significance of the results depended onthe soil type (Ma et al., 2002). Using the BCR method, organicmatter bound fraction of Cd and other heavy metals declined in thepresence of E. fetida species (Wen et al., 2004).

In the current study, the residual (HNO3-extracted or sulfide) Cdfaction was decreased in the presence of earthworm, suggestingthat ingestion of Cd-polluted soils is also an important factordetermining the concentration of Cd in the residual fraction.Similarly, the sulfide-bound fraction of Cd and other heavy metalswas declined in the presence of epigeic E. fetida species (Wen et al.,2004). In contrast, the presence of anecic L. terrestris speciesincreased both Cu and Zn bound to the sulfide fraction using theBCR method (Ruiz et al., 2011).

These conflicting results can be mainly due to differences in (i)soil characteristics, (ii) pollution mode, (iii) sequential extractionprocedures, (iv) metal type and (v) even physiological andecological differences among earthworm species. The individualeffect of earthworm activity on metal concentration and speciationis largely soil and metal specific, and could be attributed to changesin soil pH and soluble organic carbon and even the complexinteraction between earthworms and microbial community (Wenet al., 2004; Sizmur and Hodson, 2009; Sizmur et al., 2011).

Furthermore, our data showed the influence of earthwormaddition on Cd fractions was only limited to the addition of10 mg Cd kg�1 soil (Table 2), indicating the potential Cd toxicity forearthworms when added at 20mg Cd kg�1 in soil. This is supportedby a decline in earthworm dry weight with Cd addition, which wassimilar at both Cd addition rates (see Section 3.2).

3.1.3. AMF effectArbuscular mycorrhizal inoculation had a significant (P < 0.01)

influence only on the inorganic bound and residual (HNO3-extrac-ted) Cd fractions (Table 1). The present results showed that inocu-lation of G. intraradices and G. mosseae species resulted in no majorchanges in concentration of the soluble/exchangeable and organicbound Cd fractions at both levels of Cd addition, a trend opposed toearthworm effect. This emphasizes the potential capacity of AMFsymbiosis to at least not affect the mobility and bioavailability of Cdin these calcareous soils, as thewater soluble/exchangeable fractionis considered readily mobile and available to plants and other or-ganisms (Sposito et al., 1982; Jalali and Arfania, 2011). A significantdecrease in soil pH from 8.15 in non-inoculated pots to 7.65 ininoculated pots (data not shown), and a decline in DOC concentra-tion only with G. mosseae inoculation observed in our previousexperiment (Aghababaei et al., 2014) did notmodify the amounts ofCd species associated with the soluble/exchangeable and organicfractions. This suggests that changes in the soil chemical charac-teristics with AMF inoculation were not reflected at all in Cd speci-ation, especially the soluble/exchangeable Cd fraction, in themycorrhizosphere. A similar response of soluble/exchangeable Cd(Aghababaei et al., 2014) and Se (Munier-Lamy et al., 2007) fractionsto AMF inoculationwas reported. In another study, inoculationwithGlomus spp. resulted in no change or a decline in mobile forms ofother metals (Pb and Zn) in mine tailings, depending on both thelevel ofmetals in tailings and extracting agent used (Ma et al., 2006).AMF inoculation had no influence on the concentration of Cd asso-ciated with organic fraction (Table 1), mostly likely due to the un-affected SOM content by the same AMF species (Aghababaei et al.,2014). In contrast, the influence of G. mosseae inoculation on metal(Cu, Zn and Pb) speciation was seen only for the organic metalfraction, increasing the metals associated with the organic fractionusing the Tessier's procedure (Huang et al., 2005).

A decline in the inorganic bound Cd fraction due to AMF sym-biosis (Fig. 1) was accompanied by an increase in its residual frac-tion (Table 3) although the relative contribution of both fractions to

the total Cd recovered did not vary with AMF inoculation (Fig. 1 andTable 3). This difference is probably due to the Cd recovery values oflower than 100%. This finding indicates that AMF symbiosis wouldprobably transform the inorganic bound (less mobile and available)into residual (immobile and unavailable) fractions; again contraryto the effect observed with earthworm addition. The decreasedinorganic bound fraction by AMF inoculation might be due to theincreased immobilization of Cd from this fraction to residual formsfollowing the solubilization of inorganic precipitates by microbialand enzyme activities (Subramanian et al., 2009; Aghababaei et al.,2014). Mycorrhizal fungi can stimulate microbial activity(Aghababaei et al., 2014) and acidify the rhizosphere by releasingsimple organic acids such as citric and oxalic acids (Leyval andJoner, 2001). The release of carboxylic acids by AMF may solubi-lize carbonate and (hydro)oxide bound heavymetals (Giasson et al.,2005), probably increasing their accumulation in the residualfraction. The mycorrhizal fungi have also the potential to solubilizeother metals (Zn, Cu and Pb) from the less soluble forms (Fominaet al., 2005).

The residual Cd fraction measured in the current study repre-sents largely the sulfide forms (non-silicate-bound), since heavymetals in sulfide forms are generally extracted with the help of 4 MHNO3 as the last step in the sequential extraction procedure(Sposito et al., 1982). Furthermore, the metals in the crystallinelattice of soil silicates would not be expected to change in the short-term. Assuming the residual fraction specifically represents the Cdassociated with sulfide minerals, the increased sulfide fractionmight be due to high Cd affinity and binding capacity by fungalstructures in either intraradical or extraradical hyphae (Joner et al.,2000; Gonz�alez-Ch�avez et al., 2006; Janou�skov�a and Pavlíkov�a,2010). Heavy metals may be bound nonspecifically to thiolscysteine SH-groups of fungal cell wall proteins or to low molecularcarrier-proteins rich in cysteine, the so-called metallothioneins(Gonz�alez-Ch�avez et al., 2006; Chakraborty and Roy, 2006). Theincreased sulfide Cd fraction could also be due to the formation ofgrains and aggregates of metal sulfides in mycorrhizal networks(Cabala et al., 2009). By using electron scanning microscopymethods, Cabala et al. (2009) confirmed the formation of secondaryphases during biological interactions of mycorrhizae in mining andsmelting of PbeZn ores in southern Poland. They reported that themycorrhizal communities identified in the rhizosphere were rich inZn, Pb and Fe sulfides. Nevertheless, the underlying mechanism forthe increased Cd in sulfide forms with AMF inoculation is notexactly known, and cannot be explained with this data. However,treatment with 4 M HNO3 is hardly specific to estimate only sulfideforms andmay also dissolve some other metal compounds (Kabata-Pendias, 2001; Rao et al., 2008). On the other hand, a low redoxpotential (i.e., reducing conditions) is often needed for the forma-tion of HS- and the subsequent precipitation of sulfide-metal solidsout of the solution, since sulfides of heavy metals are not commonin soils, especially in soils with good drainage (Kabata-Pendias,2001).

Our findings are therefore in accordance with an assumptionthat mycorrhizae may enhance the tolerance of host plants to Cdtoxicity by lowering its uptake into plant components (Shen et al.,2006; Janou�skov�a and Pavlíkov�a, 2010); likely due to Cd immobi-lization through its transformation from the inorganic precipitatesinto the sulfide forms. However, sequential extraction methodsshould be used in conjunction with plant bioassays to determineindirectly how AMF affect plant Cd uptake and tolerance to Cdtoxicity. This is important to establish a relationship between Cdbioavailability and chemical forms obtained by a sequentialextraction procedure and plant response.

The findings of this study do not support previous findings thatmycorrhizal inoculation can affect all the chemical fractions of Cd

Fig. 1. Effects of earthworm (NE ¼ no earthworm, WE ¼ with earthworm) and AMF inoculation (NM ¼ without inoculation, Glomus intraradices and Glomus mosseae species) on theconcentration (A) and distribution (B) of inorganic Cd fraction (mean ± standard error) in contaminated soils cultivated with maize under greenhouse conditions. Mean values(n ¼ 9) with the same letter are not significantly different (P > 0.05) between treatments by the Tukey test.

F. Aghababaei et al. / Soil Biology & Biochemistry 78 (2014) 21e29 27

(Huang et al., 2008) and other heavy metals such as Zn(Subramanian et al., 2009), Cu (Huang et al., 2008) and As (Bai et al.,2008). The inconsistency may primarily be due to the differences insequential extraction methods, which use various extractants withdifferent selectivities and specificities during the extraction pro-cesses (Ponizovsky andMironenko, 2001; Rao et al., 2008). This canmake comparisons of heavy metal distribution among solid-phasefractions obtained by different sequential methods more difficult.The impact of AMF on metal availability is also fungal species andmetal specific, and may not show a general tendency (Weissenhornet al., 1995). The effect of AMF on Cd fractionation is not still evidentdue to differences in experimental conditions used in different AMFstudies (Leyval, 2005); and certainlymuchmore studies are neededto understand changes in heavy metal fractionation in relation toAMF symbiosis.

3.2. Cd, earthworm and AMF interactive effects on soil Cd fractions

We observed that the two-way interaction effect betweenearthworm and AMF on the concentrations of soluble/exchange-able, organic bound and residual Cd fractions was not statisticallysignificant (P > 0.05) regardless of Cd addition rate (Table 1), sug-gesting that the combined influences of both soil organisms onthese fractions of soil Cd were largely independent or additive.However, the two-way interaction between earthworm and AMFwas significant (P< 0.01) for the inorganic bound fractionwhile thisinteraction did not depend on the level of Cd addition (Table 1).AMF inoculation was found to decrease the inorganic bound frac-tion only in the absence of earthworms (Fig. 1). This indicates thatthe combined impacts of earthworm and AMF on the Cd associatedwith inorganic precipitates differ from their main effects, and thattheir influences on this fraction of soil Cd aremainly dependent, notadditive. We found that the interaction effect among Cd, earth-worm and AMF treatments was statistically significant (P < 0.01)only for the residual Cd fraction (Table 1). Both AMF speciesincreased the residual Cd fraction only at high Cd levels whenearthworms were absent (Table 3), reflecting that the presence ofboth earthworms and AMF may result in an antagonistic interac-tion on this Cd fraction. Our study demonstrates that the decreasedinorganic and increased residual Cd fractions with AMF inoculationwould be potentially offset by earthworms, probably because oftheir regular movement and soil disturbance. Mechanical soildisturbance by earthworm burrowing and feeding activity has beendocumented to disconnect and damage the fungal external myce-lium (Milleret et al., 2009) and thus to prevent the developmentand establishment of a mycorrhizal mycelium network. We foundevidence for the important role of earthworm activity in reducingthe residual Cd fraction only in the presence of AMF, most likelybecause of the trophic interaction between the two organisms, in

particular earthworm selective feeding on fungal hyphae andspores (Lawrence et al., 2003; Curry and Schmidt, 2007).

3.3. Cd effects on earthworm growth and Cd bioaccumulation

Earthworm dry weight significantly decreased (26e39%) withCd addition, but decreases were similar at both Cd levels (Fig. 2).Our data shows that while earthworm dry weight leveled off athigh Cd levels, Cd uptake into the body of earthworms increasedconsistently (up to 74 mg pot�1) with Cd addition and was signifi-cantly greater at high (34 mg pot�1) than low (74 mg pot�1) Cd levels(Fig. 2). This indicates that Cd concentration in earthworms wasalso greater at high than low Cd levels, because earthworm biomasswas similar at both Cd levels. However, bioaccumulation factor(BAF) for the accumulation of Cd in earthworm body was lower athigh than low Cd levels, indicating the low transfer of Cd from thesoil environment to earthworm tissues at higher exposure levels. Ithas been reported that BAF for metals generally decreases at higherexposure levels (Van Gestel et al., 2011). This study suggested Cduptake by earthworms increased with increasing its concentrationin soil, but that its bioaccumulation factor in general was lower athigher exposure levels. This implies that high Cd level is toxic toearthworm and that this metal is significantly bioaccumulated byearthworms only at low levels. It is possible that earthworms maytake up soil Cd from the soluble/exchangeable and organicallybound fractions rather than the inorganically bound one, since thelatter fraction was higher at high than low Cd levels, while theformer fractions were similar at both Cd levels (Table 2). Soilorganic matter is a major food resource for earthworms andconsequently Cd bound to organic particles is absorbed after thetransit through their gut (Nannoni et al., 2011). A significant rela-tionship between Cd concentrations in earthworms (Allolobophorarosea and Nicodrilus caliginosus species) and its contents in theextractable, reducible and oxidizable fractions was reported byNannoni et al. (2011). Another explanation is that at high levels ofpollution earthworms may eliminate the excess of toxic metalsafter exposure, possibly by excretion of these metals (Dai et al.,2004). Previous studies have reported that earthworms L. rubellusand other species can bio-accumulate Cd and other heavy metals intheir tissues when ingesting and digesting soil materials andpassing them through their guts or through direct dermal contactwith heavy metals in the soil solution (Dai et al., 2004; Nannoniet al., 2011; Van Gestel et al., 2011). Dai et al. (2004) reported agreater BAF in L. rubellus for Cd than for other metals, and BAFvalues for this species varied from 3.6 to 6.3 in soils polluted bywastes from a metallurgic industry. Nannoni et al. (2011) observedthat Cd and Zn were the only metals bio-accumulated by twoearthworm species (A. rosea and N. caliginosus) from a mining andindustrial area. In addition, our data demonstrated the lack of AMFinfluence on uptake and bioaccumulation of soil Cd by earthworms

Fig. 2. Effects of Cd level on earthworm dry weight (A), earthworm Cd uptake (B) and earthworm Cd bioaccumulation factor (BAF) (C) in contaminated soils cultivated with maizeunder greenhouse conditions. Mean values ± standard error (n ¼ 18) with the same letter are not significantly different (P > 0.05) between treatments by the Tukey test.

F. Aghababaei et al. / Soil Biology & Biochemistry 78 (2014) 21e2928

(Table 1), suggesting additive or no interaction between earth-worms and AMF. This would also mean that earthworms may notgraze and feed on AMF hyphae. This is supported by our previousresults (Aghababaei et al., 2014) that maize root colonization byG. mosseae species increased while remained unaffected byG. intraradices species in the presence of L. rubellus activity.

4. Summary and conclusions

To our knowledge, this is the first report that indicates theinteractioneffects of earthwormsandAMFonCd fractionation in themycorrhizosphere. The current study showed main but not inter-active effects of earthworms (L. rubellus) and AMF (G. intraradicesand G. mosseae) on Cd speciation in a calcareous soil spikedwith Cd.Earthworm activity increased Cd concentrations of the non-residualfraction while reducing the residual Cd species. The only effect ofAMF inoculation on Cd speciation was a decrease in the inorganicbound Cd fraction with concomitant increases in the residual Cdfraction. Consequently, earthworm addition changed Cd availabilityand speciation more than AMF inoculation. The influences of bothsoil organisms on the non-residual fractions largely were additive,except for the Cd associatedwith the inorganic precipitates, and theresidual Cd fraction, which was additive at low Cd level but antag-onistic at high Cd level. There was no strong evidence for the inter-active effects of earthworms and AMF on easily available (i.e., thesoluble þ exchangeable species) and potentially available (i.e.,bound to organic matter) Cd fractions. Results demonstrate that theinteraction between earthworms and AMFare not interesting for Cdphyto-remediation processes in the studied environment due toindependent mechanisms by which the two organisms affect Cdfractionation. However,we recommend that alongwith a sequentialfractionation scheme, plant bioassay should also be carried out forfinding a link between soil Cd fractions and plant uptake in thepresence of both earthworm and AMF. In summary, earthwormactivity modified AMF effects on the inorganic and residual Cdfractions without an interaction effect on the soluble/exchangeableand organic bound fractions, rejecting our hypothesis that both AMFand earthworm would interactively influence Cd availability andspeciation. Practically, this study suggests that the interaction be-tween the two soil organisms in relation to the non-residual Cdfraction and its mobility likely is less important, and that only themain effect of earthworms on Cd availability could be striking in thecalcareous soils polluted with Cd.

Acknowledgments

Thanks are due to Shahrekord University for providing thefinancial support of thework reported in this paper (Grant number:122/6802). The authors would like to thank the Editor-in-Chief andreviewers for constructive comments on the manuscript.

References

Adriano, D.C., 2001. Trace Elements in the Terrestrial Environment: Biogeochem-istry, Bioavailability, and Risks of Metals, second ed. Springer-Verlag, New York.

Aghababaei, F., Raiesi, F., Hosseinpur, A., 2014. The combined effects of earthwormsand arbuscular mycorrhizal fungi on microbial biomass and enzyme activitiesin a calcareous soil spiked with cadmium. Appl. Soil Ecol. 75, 33e42.

Ali, H., Khan, E., Sajad, M.A., 2013. Phytoremediation of heavy metals e conceptsand applications. Chemosphere 91, 869e881.

Amir, S., Hafidi, M., Merlina, G., Revel, J.C., 2005. Sequential extraction of heavymetals during composting of sewage sludge. Chemosphere 59, 801e810.

Antoniadis, V., Alloway, B.J., 2002. The role of dissolved organic carbon in themobility of Cd, Ni and Zn in sewage sludge-amended soils. Environ. Pollut. 117,515e521.

Bai, J., Lin, X., Zhang, H., Junhua, W., Xueming, C., Yongming, L., 2008. The influenceof arbuscular mycorrhizal fungi on As and P uptake by maize (Zea mays L.) fromAs-contaminated soils. Appl. Soil Ecol. 38, 137e145.

Binet, F., Fayolle, L., Pussard, M., 1998. Significance of earthworms in stimulating soilmicrobial activity. Biol. Fertil. Soil 27, 79e84.

Bolan, N.S., Adriano, D.C., Duraisamy, A., Mani, P.A., Arulmozhiselvan,Duraisamy, A.K., 2003. Immobilization and phytoavailability of cadmium invariable charge soils. II. Effect of lime addition. Plant Soil 251, 187e198.Immobilization and phytoavailability of cadmium in variable charge soils. I.Effect of phosphate addition. Plant Soil 250, 83e94.

Brinza, L., Schofield, P.F., Mosselmans, J.F.W., Donner, E., Lombi, E., Paterson, D.,Hodson, M.E., 2014. Can earthworm-secreted calcium carbonate immobilise Znin contaminated soils? Soil Biol. Biochem. 74, 1e10.

Brown, G.G., Barois, I., Lavelle, P., 2000. Regulation of soil organic matter dynamicsand microbial activity in the drilosphere and the role of interactions with otheredaphic functional domains. Eur. J. Soil Biol. 36, 177e198.

Cabala, J., Krupa, P., Misz-Kennan, M., 2009. Heavy metals in mycorrhizal rhizo-spheres contaminated by ZnePb mining and smelting around Olkusz insouthern Poland. Water Air Soil Pollut. 199, 139e149.

Caravaca, F., Rold�an, A., 2003. Assessing changes in physical and biological prop-erties in a soil contaminated by oil sludges under semi-arid Mediterraneanconditions. Geoderma 117, 53e61.

Chakraborty, R., Roy, R., 2006. Microbial genomics as an integrated tool for devel-oping biosensors for toxic trace elements in the environment. In:Prasad, M.N.V., Naidu, R., Sajwan, K.S. (Eds.), Trace Elements in the Environ-ment: Biogeochemistry, Biotechnology, and Bioremediation. Taylor & FrancisGroup, Boca Raton, pp. 169e210.

Cheng, J., Wong, M.H., 2002. Effects of earthworms on Zn fractionation in soils. Biol.Fertil. Soil 36, 72e78.

Curry, J.P., Schmidt, O., 2007. The feeding ecology of earthworms: a review. Pedo-biologia 50, 463e477.

Dai, J., Becquer, T., Rouiller, J.H., Reversat, G., Bernhard-Reversat, F., Nahmani, J.,Lavelle, P., 2004. Heavy metal accumulation by two earthworm species and itsrelationship to total and DTPA-extractable metals in soils. Soil Biol. Biochem. 36,91e98.

Edwards, C.A., Bohlen, P.J., 1996. Biology and Ecology of Earthworms, third ed.Chapman and Hall, UK.

Fomina, M.A., Alexander, I.J., Colpaert, J.V., Gadd, G.M., 2005. Solubilization of toxicmetal minerals and metal tolerance of mycorrhizal fungi. Soil Biol. Biochem 37,851e866.

Giasson, P., Jaouich, A., Gagn�e, S., Moutoglis, P., 2005. Arbuscular mycorrhizal fungiinvolvement in zinc and cadmium speciation change and phytoaccumulation.Remediat. J. 15, 75e81.

Gonz�alez-Ch�avez, M.C., Carrillo-Gonz�alez, R., Wright, S.F., Nichols, K.A., 2004. Therole of glomalin, a protein produced by arbuscular mycorrhizal fungi, insequestering potentially toxic elements. Environ. Pollut. 130, 317e323.

Gonz�alez-Ch�avez, M.C., Vangronsveld, J., Colpaert, J., Leyval, C., 2006. Arbuscularmycorrhizal fungi and heavy metals: tolerance mechanisms and potential usein bioremediation. In: Prasad, M.N.V., Naidu, R., Sajwan, K.S. (Eds.), Trace Ele-ments in the Environment: Biogeochemistry, Biotechnology, and Bioremedia-tion. Taylor & Francis Group, Boca Raton, pp. 211e234.

F. Aghababaei et al. / Soil Biology & Biochemistry 78 (2014) 21e29 29

Hua, J.F., Lin, X.G., Bai, J.F., Shao, Y.F., Yin, R., Jiang, Q., 2010. Effects of arbuscularmycorrhizal fungi and earthworm on nematode communities and arsenic up-take by maize in arsenic contaminated soils. Pedosphere 20, 163e173.

Huang, Y., Li, T., Huag, Z.J., Fei, Y.H., 2008. Ectomycorrhizal fungus-induced changesof Cu and Cd speciation in the rhizosphere of Chinese pine seedlings. Pedo-sphere 18, 758e765.

Huang, Y., Tao, S., Chen, Y., 2005. The role of arbuscular mycorrhiza on change ofheavy metal speciation in rhizosphere of maize in wastewater irrigated agri-culture soil. J. Environ. Sci. 17, 276e280.

Jalali, M., Arfania, H., 2011. Distribution and fractionation of cadmium, copper, lead,nickel, and zinc in a calcareous sandy soil receiving municipal solid waste.Environ. Monit. Assess. 173, 241e250.

Janou�skov�a, M., Pavlíkov�a, D., 2010. Cadmium immobilization in the rhizosphere ofarbuscular mycorrhizal plants by the fungal extraradical mycelium. Plant Soil332, 511e520.

Joner, E.J., Briones, R., Leyval, C., 2000. Metal-binding capacity of arbuscularmycorrhizal mycelium. Plant Soil 226, 227e234.

Kabata-Pendias, A., 2001. Trace Elements in Soils and Plants. CRC Press, Boca Raton.Kirkham, M.B., 2006. Cadmium in plants on polluted soils: effects of soil factors,

hyperaccumulation, and amendments. Geoderma 137, 19e32.Kizilkaya, R., 2004. Cu and Zn accumulation in earthworm Lumbricus terrestris L. in

sewage sludge amended soil and fractions of Cu and Zn in casts and sur-rounding soil. Ecol. Eng. 22, 141e151.

Lawrence, B., Fisk, M.C., Fahey, T.J., Su�arez, E.R., 2003. Influence of nonnativeearthworms on mycorrhizal colonization of sugar maple (Acer saccharum). NewPhytol. 157, 145e153.

Leyval, C., 2005. Effect of arbuscular mycorrhizal (AM) fungi on heavy metal andradionuclide transfer to plants. In: Huang, P.M., Gobran, G.R. (Eds.), Biogeo-chemistry of Trace Elements in the Rhizosphere. Elsevier B.V., pp. 419e429.

Leyval, C., Joner, E.J., 2001. Bioavailability of heavy metals in the mycorrhizosphere.In: Gobran, G.R., Wenzel, W., Lombi, E. (Eds.), Trace Elements in the Rhizo-sphere. CRC, Boca Raton, pp. 165e185.

Liu, L., Chen, H., Cai, P., Liang, W., Huang, Q., 2009. Immobilization and phytotoxicityof Cd in contaminated soil amended with chicken manure compost. J. Hazard.Mater. 163, 563e567.

Lukkari, T., Teno, S., Vaeisaenen, A., Haimi, J., 2006. Effects of earthworms ondecomposition andmetal availability in contaminated soil: microcosm studies ofpopulations with different exposure histories. Soil Biol. Biochem. 38, 359e370.

Ma, Y., Dickinson, N.M., Wong, M.H., 2006. Beneficial effects of earthworms andarbuscular mycorrhizal fungi on establishment of leguminous trees on Pb/Znmine tailings. Soil Biol. Biochem. 38, 1403e1412.

Ma, Y., Dickinson, N.M., Wong, M.H., 2002. Toxicity of Pb/Zn mine tailings to theearthworm Pheretima and the effects of burrowing on metal availability. Biol.Fertil. Soil 36, 79e86.

Ma, L.Q., Rao, G.N., 1997. Chemical fractionation of cadmium, copper, nickel, andzinc in contaminated soils. J. Environ. Qual. 26, 259e264.

Milleret, R., Le Bayon, R.C., Gobat, J.M., 2009. Root, mycorrhizal and earthworminteractions: their effects on soil structuring processes, plant and soil nutrientconcentration and plant biomass. Plant Soil 316, 1e12.

Moral, R., Gilkes, R.J., Jord�An, M.M., 2005. Distribution of heavy metals in calcareousand non-Calcareous soils in Spain. Water Air Soil Pollut. 162, 127e142.

Munier-Lamy, C., Deneux-Mustin, S., Mustin, C., Merlet, D., Berthelin, J., Leyval, C.,2007. Selenium bioavailability and uptake as affected by four different plants ina loamy clay soil with particular attention to mycorrhizae inoculated ryegrass.J. Environ. Radioact. 97, 148e158.

Nannoni, F., Protano, G., Riccobono, F., 2011. Fractionation and geochemical mobilityof heavy elements in soils of a mining area in northern Kosovo. Geoderma 161,63e73.

Natal-da-Luz, T., Ojed, G., Pratas, J., vanGestel, C.A.M., Sousa, J.P., 2011. Toxicity toEisenia andrei and Folsomia candida of a metal mixture applied to soil directly orvia an organic matrix. Ecotoxicol. Environ. Saf. 74, 1715e1720.

Pelfrene, A., Waterlot, C., Mazzuca, M., Nisse, C., Cuny, D., Richard, A., Denys, S.,Heyman, C., Roussel, H., Bidar, G., Douay, F., 2012. Bioaccessibility of trace ele-ments as affected by soil parameters in smelter-contaminated agricultural soils:a statistical modeling approach. Environ. Pollut. 160, 130e138.

Ponizovsky, A., Mironenko, E., 2001. Speciation and sorption of lead (II) in soils. In:Iskandar, I.K., Kirkham, M.B. (Eds.), Trace Elements in Soil: Bioavailability, Flux,and Transfer. Lewis Publishers, Boca Raton, pp. 261e279.

Rao, C.R.M., Sahuquillo, A., Lopez Sanchez, J.F.L., 2008. A review of the differentmethods applied in environmental geochemistry for single and sequential

extraction of trace elements in soils and related materials. Water Air Soil Pollut.189, 291e333.

Renella, G., Adamo, P., Bianco, M.R., Landi, L., Violante, P., Nannipieri, P., 2004.Availability and speciation of cadmium added to a calcareous soil under variousmanagements. Eur. J. Soil Sci. 55, 123e133.

Ruiz, E., Alonso-Azc�arate, J., Rodríguez, L., 2011. Lumbricus terrestris L. activity in-creases the availability of metals and their accumulation in maize and barley.Environ. Pollut. 159, 722e728.

Ruiz, E., Rodríguez, L., Alonso-Azc�arate, J., 2009. Effects of earthworms on metaluptake of heavy metals from polluted mine soils by different crop plants.Chemosphere 75, 1035e1041.

Schroeder, M.S., Janos, D.P., 2004. Phosphorus and intraspecific density alter plantresponses to arbuscular mycorrhizas. Plant Soil 264, 335e348.

Shen, H., Christie, P., Li, X., 2006. Uptake of zinc, cadmium and phosphorus byarbuscular mycorrhizal maize (Zea mays L.) from a low available phosphoruscalcareous soil spiked with zinc and cadmium. Environ. Geochem. Health 28,111e119.

Sizmur, T., Palumbo-Roe, B., Watts, M.J., Hodson, M.E., 2011. Impact of the earth-worm Lumbricus terrestris (L.) on As, Cu, Pb and Zn mobility and speciation incontaminated soils. Environ. Pollut. 159, 742e748.

Sizmur, T., Hodson, M.E., 2009. Do earthworms impact metal mobility and avail-ability in soil? e a review. Environ. Pollut. 157, 1981e1989.

Smith, S.R., 2009. A critical review of the bioavailability and impacts of heavy metalsin municipal solid waste composts compared to sewage sludge. Environ. Int. 35,142e156.

Soil Survey Staff, 2010. Keys to Soil Taxonomy, 11th ed. USDA-Natural ResourcesConservation Service, Washington, DC.

Sposito, G., Lund, L.J., Chang, A.C., 1982. Trace metal chemistry in arid-zone fieldsoils amended with sewage sludge: I. Fractionation of Ni, Cu, Zn, Cd and Pb insoil phases. Soil Sci. Soc. Am. J. 46, 260e264.

Strobel, B.W., Borggaard, O.K., Hansen, H.C.B., Andersen, M.K., Raulund-Rasmussen, K., 2005. Dissolved organic carbon and decreasing pH mobilizecadmium and copper in soil. Eur. J. Soil Sci. 56, 189e196.

Subramanian, K.S., Tenshia, V., Jayalakshmi, K., Ramachandran, V., 2009. Biochem-ical changes and zinc fractions in arbuscular mycorrhizal fungus (Glomusintraradices) inoculated and uninoculated soils under differential zinc fertil-ization. Appl. Soil Ecol. 43, 32e39.

Tao, S., Chen, Y.J., Xu, F.L., Cao, J., Li, B.G., 2003. Changes of copper speciation inmaize rhizosphere soil. Environ. Pollut. 122, 447e454.

Udovic, M., Lestan, D., 2010. Fractionation and bioavailability of Cu in soil reme-diated by EDTA leaching and processed by earthworms (Lumbricus terrestris L.).Environ. Sci. Pollut. Res. 17, 561e570.

Udovic, M., Lestan, D., 2007. The effect of earthworms on the fractionation andbioavailability of heavy metals before and after soil remediation. Environ. Pol-lut. 148, 663e668.

Van Gestel, C.A.M., Ortiz, M.D., Borgman, E., Verweij, R.A., 2011. The bio-accumulation of molybdenum in the earthworm Eisenia andrei: influence of soilproperties and ageing. Chemosphere 82, 1614e1619.

Versteegh, E.A.A., Black, S., Hodson, M.E., 2014. Environmental controls on theproduction of calcium carbonate by earthworms. Soil Biol. Biochem. 70,159e161.

Vig, K., Megharaj, M., Sethunathan, N., Naidu, R., 2003. Bioavailability and toxicity ofcadmium to microorganisms and their activities in soil: a review. Adv. Environ.Res. 8, 121e135.

Weissenhorn, I., Leyval, C., Berthelin, J., 1995. Bioavailability of heavy metals andabundance of arbuscular mycorrhiza in a soil polluted by atmospheric depo-sition from a smelter. Biol. Fertil. Soil 19, 22e28.

Wen, B., Hu, X., Liu, Y., Wang, W., Feng, M., Shan, X., 2004. The role of earthworms(Eisenia fetida) in influencing bioavailability of heavy metals in soils. Biol. Fertil.Soil 40, 181e187.

Wen, B., Liu, Y., Hu, X., Shan, X., 2006. Effect of earthworms (Eisenia fetida) on thefractionation and bioavailability of rare earth elements in nine Chinese soils.Chemosphere 63, 1179e1186.

Wong, J.W.C., Li, K.L., Zhou, L.X., Selvam, A., 2007. The sorption of Cd and Zn bydifferent soils in the presence of dissolved organic matter from sludge. Geo-derma 137, 310e317.

Yu, X., Cheng, J., Wong, M.H., 2005. Earthwormemycorrhiza interaction on Cd up-take and growth of ryegrass. Soil Biol. Biochem. 37, 195e201.