Heavy metal-immobilizing organoclay facilitates polycyclic aromatic hydrocarbon biodegradation in mixed-contaminated soil

Journal of Hazardous Materials 298 (2015) 129–137

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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat

Heavy metal-immobilizing organoclay facilitates polycyclic aromatic hydrocarbon biodegradation in mixed-contaminated soil Bhabananda Biswas a,b , Binoy Sarkar a,b , Asit Mandal a,c , Ravi Naidu a,b,∗,1 a b c

Centre for Environmental Risk Assessment and Remediation, University of South Australia, Mawson Lakes Campus, SA 5095, Australia Cooperative Research Centre for Contamination Assessment and Remediation of the Environment, P.O. Box 486, Salisbury, SA 5106, Australia Division of Soil Biology, Indian Institute of Soil Science, Bhopal, Madhya Pradesh, India

h i g h l i g h t s

g r a p h i c a l

a b s t r a c t

• A novel metal-immobilizing organ• • • •

oclay (MIOC) synthesized and characterized. MIOC immobilizes toxic metals and reduces metal bioavailability. It enhances PAH-bioavailability to soil bacteria. It improves microbial growth and activities in mixed-contaminated soils. MIOC facilitates PAH-biodegradation in metal co-contaminated soils.

a r t i c l e

i n f o

Article history: Received 7 February 2015 Received in revised form 5 May 2015 Accepted 8 May 2015 Available online 12 May 2015 Keywords: Metal-immobilizing organoclay Mixed contaminants Microbial activities Bioavailability PAH-biodegradation

a b s t r a c t Soils contaminated with a mixture of heavy metals and polycyclic aromatic hydrocarbons (PAHs) pose toxic metal stress to native PAH-degrading microorganisms. Adsorbents such as clay and modified clay minerals can bind the metal and reduce its toxicity to microorganisms. However, in a mixedcontaminated soil, an adsorption process more specific to the metals without affecting the bioavailability of PAHs is desired for effective degradation. Furthermore, the adsorbent should enhance the viability of PAH-degrading microorganisms. A metal-immobilizing organoclay (Arquad® 2HT-75-bentonite treated with palmitic acid) (MIOC) able to reduce metal (cadmium (Cd)) toxicity and enhance PAH (phenanthrene) biodegradation was developed and characterized in this study. The MIOC differed considerably from the parent clay in terms of its ability to reduce metal toxicity (MIOC > unmodified bentonite > Arquad–bentonite). The MIOC variably increased the microbial count (10–43%) as well as activities (respiration 3–44%; enzymatic activities up to 68%), and simultaneously maintained phenanthrene in bioavailable form in a Cd-phenanthrene mixed-contaminated soil over a 21-day incubation period. This study may lead to a new MIOC-assisted bioremediation technique for PAHs in mixedcontaminated soils. © 2015 Published by Elsevier B.V.

∗ Corresponding author. Tel.: +61 8 8302 5041; fax: +61 8 8302 3124. 1 Present address: Professor, Faculty of Science and Information Technology, University of Newcastle, Callaghan, NSW 2308, Australia. http://dx.doi.org/10.1016/j.jhazmat.2015.05.009 0304-3894/© 2015 Published by Elsevier B.V.

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1. Introduction Organic contaminants such as polycyclic aromatic hydrocarbons (PAHs) can adversely impact environmental quality and human health, and greater dangers can emerge when these contaminants co-exist with heavy metals. PAH-contaminated soils often contain high concentrations of metals (e.g., cadmium (Cd)) also originating from similar sources (natural and anthropogenic) as those of PAHs [1,2]. Gasworks, motor vehicle emissions and smelting are the major sources of these mixed-contaminants [3]. Phosphate fertilisers are also a significant source of Cd in agricultural soils [4]. Bioremediation is a cost-effective method for removing contaminants compared to many other physical and chemical techniques [5]. However, remediation of PAHs from a mixed-contaminated site can be challenging in the presence of metals because the latter is likely to adversely compromise microbial activity. The evidence of Cd stress to soil microorganisms is well documented for C-mineralisation, ATP production, enzymatic function (e.g., dehydrogenase) and community shift [6]. Whilst Cd reduces microorganisms’ enzymatic activities, the presence of PAHs like phenanthrene further increases the toxicity level [7]. Any such impact on soil microbial activity may eventually inhibit the biodegradation process [8]. Adsorption/immobilization of metals potentially reduces this toxicity [9–11]. In mixed-contaminated sites, however, the adsorbent should immobilize metals, but not make PAHs unavailable to the degraders. Therefore, an adsorption process more specific to the metals without affecting the bioavailability of PAHs is desired for this remediation strategy. Furthermore, the adsorbent should be compatible with the PAH-degrading microorganisms in order to enhance their proliferation, which would then translate into effective biodegradation [12]. Synthetic chelates, for example ethylenediaminetetraacetic acid (EDTA) and ethylenediamine-N,N´ı-disuccinic acid (EDDS), increase metal mobility in soils for facilitating phytoremediation or soilwashing [13]. However, a chelating agent when grafted on a clay mineral can adsorb/immobilize heavy metals. The above aminopolycarboxylic acid-type compounds, however, are incompatible with microorganisms [14]. Conversely, naturally occurring long chain fatty acid chelates, e.g., palmitic acid (PA) and stearic acid (SA), are non-toxic and compatible with soil microorganisms [15]. Therefore, adsorption of metals on PA/SA-grafted clays could reduce metal bioavailability/toxicity and improve PAH-degrading bacterial proliferation in mixed-contaminated scenario [16]. Clay minerals are ubiquitous and non-toxic in the natural environment. A metal-immobilising organoclay (MIOC) can be prepared by grafting PA/SA on clay minerals. Organoclay is initially prepared using a suitable surfactant and then the chelating agent is grafted onto the organoclay [17]. Since both the host clay and chelate are non-toxic, the resulting product presumably becomes compatible with the PAH-degrading microorganisms. However, biodegradation efficiency would depend on the clay types, nature of modifying surfactant and target contaminants. The surfactants themselves vary in their toxicity to microorganisms depending on their molecular structure. For example, di(hydrogenated tallow) dimethylammonium (Arquad® 2HT-75) is significantly less toxic to soil microbial activities than other commonly used surfactants like hexadecyl trimethylammonium (HDTMA) and octadecyl trimethylammonium (ODTMA) [18,19]. The use of biocompatible Arquad-organoclay for preparing a MIOC and its compatibility to microorganisms in a mixedcontaminated soil was never studied previously. Malakul et al. [20] used cetylbenzyl dimethylammonium (CBDA) to prepare an organoclay and grafted PA/SA onto it. Their MIOC reduced Cd and Pb toxicity to naphthalene-degrading bacterium, Pseudomonas putida,

in an aqueous suspension. While previous studies assessed microbial compatibility of MIOC in aqueous suspension using a single bacterium [21,22], a similar assessment in natural soils representing the total microbial activity was not attempted earlier. Since microbial consortia may play a more important role than a single microorganism during PAH-biodegradation [23], it is necessary to assess the effect of MIOCs on native microbial consortia instead of a single degrader. In order to address PAH-biodegradation in mixed-contaminated sites, we have developed a MIOC by grafting PA on Arquadorganobentonite and explored the mechanism of its compatibility with native microorganisms in a Cd-phenanthrene mixedcontaminated soil. 2. Materials and methods 2.1. Material preparation A Watheroo bentonite (purchased from Bentonite Products Western Australia; cation exchange capacity (CEC): 85.8 cmol [p+ ] kg−1 ; conductivity: 3.74 dSm−1 ; clay: 83.4%, silt: 2.5%, fine sand: 14.1%; particle size: ≤75 ␮m; mineralogical properties: major mineral is montmorillonite mixed with traces of quartz, dolomite and halite; chemical composition: main compounds are 47.06% SiO2 , 25.28% Al2 O3 , 6.51% TiO2 , 2.34% Na2 O, 2.81% Fe2 O3 , 0.67% CaO, 0.23% K2 O) was used as the starting material. For MIOC preparation, a cationic surfactant, Arquad® 2HT-75 (Sigma–Aldrich) and a chelating agent, palmitic acid (PA) (>98% pure, Sigma–Aldrich), were utilized. A two-step treatment was applied. First, bentonite (B) was treated with Arquad® 2HT-75 at a loading rate of 100% of CEC of the clay [24]. The organoclay was termed AB (Arquad–bentonite). Then, AB was added to ethanol-dissolved PA (10.99 g PA in 1000 mL ethanol-water mixture (1:1)). The quantity of PA was equivalent to 100% CEC of the clay. The treatment medium (pH 8–8.5) was gently stirred for 4 h on a magnetic stirrer (Model: IKA® C-MAG HS-7, Malaysia). This elevated pH maintained carboxylate form of PA and allowed it to bind to AB through mixed bilayer formation [17]. The product was collected by centrifugation at 3400 g-force for 20 min and stored after extensive washing with Milli-Q water and drying at 60 ◦ C. The product was termed ABP (Arquad–bentonite–PA). 2.2. Material characterization 2.2.1. X-ray diffraction (XRD) The powdered adsorbents (B, AB and ABP) were pressed in stainless steel sample holders and XRD patterns were obtained using CuK␣ radiation ( = 1.540598 Å) on a PANalytical Empyrean X-ray diffractometer equipped with PIXcel3D detector (PANalytical Inc., The Netherlands) operating at 40 kV and 40 mA between 2◦ and 50◦ 2 at a step size of 0.0263◦ . The basal spacing (d) was calculated from the 2 value using Bragg’s equation (n = 2d sin, where, n = an integer, ␭ = wavelength and  = the scattering angle). 2.2.2. Surface area and surface charge BET (Brunauer–Emmett–Teller) specific surface area (SSA) of the adsorbents was measured employing a Gemini 2380 surface area analyzer (Micrometrics, USA) by measuring N2 adsorption at liquid nitrogen temperature. Zeta potential values of a 0.05% suspension (w/v) of B, AB and ABP in Milli-Q water were determined by a zeta potential analyzer (Nicomp 380 ZLS, USA). The clay suspensions’ pH values were adjusted with dilute NaOH or HCl (0.5 M) to achieve a pH range (2–10) which is often encountered under environmental conditions, and zeta potentials were measured at various pH values.

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Table 1 The characteristics of field and spiked soils before and after incubation. PAH spike

Soil

Clay amended

Spiked Cd concentration

Incubation day

No Yes

Field soil

– NC B AB ABP NC

– –

Before incubation

0 10 20 0 10 20 0 10 20 0 10 20

After incubation (day 21)

B

AB

ABP

2.3. Soil spiking and microcosm experiment A field soil (0–10 cm depth, pH 7.6, moisture = 13%, Cd = 50 ␮g kg−1 dry soil, phenanthrene below detection limit) was collected from the Mawson Lakes area in South Australia. Moist soil sieved to <2 mm was used for contaminant spiking and microbial activity experiments. A total of 1 kg fresh soil was used for the clay amendments followed by phenanthrene spiking. First, each of the three adsorbents was mixed with soil (900 g) in a 2 L glass jar to obtain a final 5% loading rate (w/w, dry-weight basis) in the entire 1 kg soil and maintained at room temperature (at 60% water holding capacity (WHC) of the mixture) for 7 days for homogenous mixing. A control soil without clay amendment (NC) was also maintained. An autoclaved soil served as a biotic control. Then, phenanthrene (200 mg kg−1 dry soil) was spiked by following the partial treatment protocol [25]. Briefly, 1.6 mL of phenanthrene stock solution (125 mg mL−1 in acetone) was mixed with the remaining soil (100 g), which gave 2000 mg kg−1 phenanthrene concentration. The spiked soil (100 g) was enclosed in a glass jar for 5 min followed by solvent evaporation in air for 16 h (to reduce the effect of solvent on microorganisms). The 100 g soil was then thoroughly mixed with the clay-amended soil (900 g) to obtain a final 200 mg kg−1 of phenanthrene and incubated for another 24 h at room temperature. The total mixing period for the clay-amended soils was 8 days (7 days for clay-soil mixture followed by one day with clay-soil-phenanthrene). The recovery of phenanthrene from the initial soil by n-butanol extraction (Section 2.5) was 78–80%. A portion (200 g) of this clay/phenanthrene-mixed soil was then used to make microcosms in 250 mL glass jars sealed with a screw cap. Different concentrations of Cd (0, 10 and 20 mg kg−1 dry soil) as Cd(NO3 )2 (Scharlab S.L) were spiked into each microcosm. After a further 24 h mixing on an end-over-end shaker, the microbial activity in the mixed-contaminated soil began counting at ‘day zero’. Two sets of microcosms were established: the first was dedicated for microbial respiration measurements; and the second was used for bacterial enumeration, soil enzymes and contaminants analyses. 2.3.1. Soil characterization Total organic carbon (TOC), total N, pH and CECe were measured in the field and spiked soils before and after incubation using standard methods [26]. The field soil fell within the textural class of a silty-clay-loam, containing 39.6% clay, 25.8% sand and 34.6% silt (Table 1). The field soil contained 0.27% N, 2.75% TOC and 28.90 cmol [p+] kg−1 CECe (Table 1). The pH of all soils fell in the neutral to slightly alkaline range (7.33–7.67) throughout the incubation period (Table 1). Attributable to the pH buffering action of clays

pH 7.65 7.64 7.75 7.56 7.67 7.54 7.33 7.35 7.60 7.59 7.63 7.45 7.32 7.32 7.64 7.59 7.47

CECe (cmol [p+ ] kg−1 ) 28.90 27.84 30.17 29.60 28.03 29.76 29.48 28.57 30.91 30.87 29.12 27.71 28.23 27.07 26.01 27.37 27.61

%TOC 2.75 2.81 2.64 4.41 4.42 2.62 2.76 2.70 2.58 2.53 2.62 4.02 4.05 4.09 4.48 4.57 4.43

± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ±

%TOC:N 0.03 0.06 0.09 0.91 0.06 0.09 0.11 0.09 0.15 0.15 0.04 0.14 0.20 0.21 0.21 0.18 0.19

10.63 10.65 10.66 14.95 14.56 10.34 10.58 10.04 10.52 10.28 10.09 13.11 12.83 12.86 15.17 15.61 15.61

± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ±

1.76 1.80 1.69 2.91 2.91 1.67 1.77 1.72 1.65 1.62 1.67 2.63 2.64 2.67 2.96 3.03 2.93

[27], the phenanthrene-spiked soil without any clay amendment produced slightly lower pHs than the clay-amended soils which otherwise maintained values closer to the field soil. 2.3.2. Microbial respiration Total soil respiration was measured by the CO2 trapping method [28]. A glass vial containing 0.5 M NaOH (10 mL) was hung inside the undisturbed and closed microcosm. The CO2 released by the microorganisms was trapped by the alkali which was collected and replaced with a fresh solution at 48 h intervals. The autoclaved soil served as the control. The amount of CO2 trapped was determined by titration with 0.1 M HCl following fixation with 1 M BaCl2 (1 mL). 2.3.3. Bacterial count (CFU) On sampling days, soil (1 g) from each treatment was dispersed (in duplicate) in sodium hexametaphosphate (35 g L−1 ) and sodium carbonate (7 g L−1 ) solution (10 mL) [29] by vigorous shaking on an orbital shaker at 300 rpm. Serially diluted suspension (dilution factor 10−6 and 10−7 ) was spread on Luria agar plates, and incubated for 3–5 days at room temperature for CFU counting. The CFU was nil for the autoclaved soil. 2.3.4. Dehydrogenase activity Dehydrogenase activity (DHA) was determined by the formation of triphenyl formazan (TPF) from 2,3,5-triphenyl–tetrazolium chloride (TTC) by soil microorganisms [30]. A fresh TTC solution (3%; 0.5 mL) was added to soil (1 g) taken from each microcosm (in duplicate) and mixed gently in a 15 mL glass vial. Sterile water (2 mL) was added on the soil surface and gentle tapping was applied to remove any bubble (gaseous oxygen) [19]. All procedures were undertaken in the laminar flow cabinet using sterile water and glassware to minimize microbial contamination. Following incubation (24 h; 37 ◦ C), TPF was extracted with methanol (10 mL) by vigorous shaking. The extract was centrifuged (3400 g-force; 20 min) to discard the soil particles. The concentration of TPF was determined in the clear supernatant at 485 nm wavelength by using a quartz microplate on a Synergy HT Multi-Mode Microplate Reader (BioTek USA). No TPF was recorded in the autoclaved soil. 2.4. Bioavailable Cd The potentially bioavailable fraction of Cd was determined using a physiologically-based extraction technique (PBET) [31]. Moist soil (0.5 g) was taken from the microcosm (in duplicate) and mixed with 0.4 M glycine (pH 2.2) at 1:100 ratio. The mixture was shaken on a vertical suspension mixer (1 h; 37 ◦ C) followed by centrifugation at 3400 g-force for 20 min. The pH of the supernatant was

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Fig. 1. X-ray diffraction patterns of bentonite (B), Arquad® 2HT-75 modified bentonite (AB) and AB-palmitic acid modified bentonite (ABP).

recorded to confirm that it had not fluctuated by more than 0.5 log units [32]. Following filtration (0.45 ␮m nylon filter), Cd was measured by inductively coupled plasma–mass spectrometry (ICP–MS) (Model 7500c, Agilent Technologies, Japan). 2.5. Bioavailable phenanthrene The bioavailable fraction of phenanthrene was measured in selective microcosms with a modified protocol of n-butanol (BuOH) extraction [33,34]. Briefly, moist soil (1 g) was taken (in duplicate) in BuOH (10 mL) into a Teflon® vial followed by mild vortex for 5 s. All the procedures were carried out twice separately using same conditions. Following centrifugation (7600 g-force; 10 min), the supernatant was filtered through 0.2 ␮m PTFE filter before measuring the phenanthrene concentration with high performance liquid chromatography (HPLC) (Model 1200, Agilent Technologies, Japan) against a high purity standard (phenanthrene analytical standard, Supelco, Sigma–Aldrich). The HPLC conditions were: Eclipse XDB-C18, 4.6 × 150 mm, 5 ␮m particle size column; 85% methanol15% water mobile phase; 1.0 mL min−1 flow rate; UV detection at 254 nm. 2.6. Statistical analysis and graphical presentation Two-way ANOVA was performed to determine the effect of adsorbent materials and Cd toxicity on the biological activity in microcosm experiments using IBM SPSS 20.0 software package. Duncan’s test at 95% confidence level was conducted to examine whether differences in effect were significant. Minitab 16 and MS Excel 10 were also used for other statistical analyses and graphical presentation of the data. 3. Results and discussion 3.1. Material characterization 3.1.1. XRD The XRD pattern showed that modifying bentonite with Arquad® 2HT-75 followed by PA had changed the original clay

structure (Fig. 1). Whilst diffraction reflections due to quartz impurity existed in all three materials, those of halite (NaCl) in bentonite (B) disappeared following modifications (AB and ABP) due to washing out of the soluble compounds (Fig. 1). The primary reflection of 0 0 1 plane in B, AB and ABP showed a gradual shift in position to the left (lesser 2␪ values), which corresponded to basal spacing values (d) of 1.27, 3.03 and 4.03 nm, respectively. The insertion of surfactant molecules (Arquad® 2HT-75) in the bentonite interlayer through cation exchange reaction might impart a paraffin-type monolayer arrangement of the guest molecules in AB resulting in a greater d value (3.03 nm) [9]. This value further increased to 4.03 nm in ABP, which indicated a successful grafting of the chelating agent (PA) in the interlayer space. The Arquad molecules existing in the AB interlayer would have attracted the PA molecules through hydrophobic interaction and facilitated the grafting. A d value of 4.03 nm in ABP indicated a flattened plate formation by Arquad and PA through interactions between surfactant alkyl chain-silicate surface and the alkyl chainalkyl chain of the surfactant and PA molecules [9,24]. Malakul et al. [17] reported an expansion of montmorillonite interlayer from 0.94 nm to 1.7 nm due to modification with CBDA and PA. The expansion observed in the current study is considerably greater than previously reported because: (a) the bentonite used in this study had a greater CEC value (85.8 cmol [p+ ] kg−1 against 77 cmol [p+ ] kg−1 ) of montmorillonite used by Malakul et al. [17]; (b) Arquad has a different molecular structure than CBDA; and (c) Arquad encouraged a more favorable grafting reaction. The degree of PA grafting in MIOC in the current study is noticeably greater than in previous reports, which was further confirmed by the appearance of diffraction peaks due to 0 0 2 and 0 0 3 planes in ABP corresponding to d values 1.99 and 1.29 nm, respectively [9,24]. 3.1.2. Surface properties The unmodified bentonite exhibited a SSA of 15.4 m2 g−1 , which decreased to 2.63 m2 g−1 due to Arquad® 2HT-75 intercalation (AB) and further dropped to 1.27 m2 g−1 following PA grafting (ABP). The decrease in SSA is attributed to the pore blocking effect through gradual incorporation of surfactant and PA molecules in bentonite interlayers [35]. These results supplement the XRD findings that

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Fig. 2. Zeta potential of bentonite (B), Arquad® 2HT-75 modified bentonite (AB) and AB-palmitic acid modified bentonite (ABP) at environmental pH values.

Arquad entering in the bentonite interlayer carries out the interspace grafting of PA in the MIOC and increases the d value (Fig. 1). The modifications also changed the surface charge properties of the clay products at a wide range of pH values (Fig. 2). The MIOC (ABP) possessed a higher net negative charge (zeta potential;  = –45 mV) compared to B ( = −25 mV) and AB ( = 14 mV) at around pH 7.65. ABP exhibited greater negative  values than B and AB over a frequently encountered environmental pH range (3–10). This is attributed to the successful grafting of PA into the organoclay structure. The negative surface charge of ABP indicated its strong affinity for Cd2+ [36], which would reduce Cd bioavailability to microorganisms during PAH biodegradation. 3.2. Microbial activities 3.2.1. Bacterial count (CFU) The CFU in phenanthrene-spiked soils varied insignificantly (p > 0.05) due to clay amendments (Fig. 3). ABP enhanced bacte-

Fig. 3. Bacterial count (CFU) in the clay-amended phenanthrene-spiked soils (7 days mixing for clay + soil and 1 day mixing for clay + soil + phenanthrene) (ns = not significant).

rial growth to the highest level (2.5 × 108 ± 0.97 × 108 ), followed by AB (2.47 × 108 ± 1.56 × 108 ) and B (0.71 × 108 ± 0.15 × 108 ) (Fig. 3). Concerning the treatments, only bentonite showed a smaller CFU than the unamended soil immediately after phenanthrene spiking. Expandable clay, especially bentonite, may induce hydrocarbondegrading bacterial growth in aqueous suspension [27,37], but complex processes interplay in soils involving many other extrinsic factors. Migration of phenanthrene and/or soil organic matter into bentonite pores might have reduced the bioavailability of carbon sources to the bacteria [38]. Ugochukwu [37] observed no effect of bentonite on bacterial growth after 48 h of incubation, but it enhanced CFU as the incubation period progressed further. In the current study bentonite enhanced bacterial growth in mixed-contaminated soils (compared to the control treatment) as the incubation time progressed. At day 1, a 20 mg kg−1 Cd con-

Fig. 4. Logarithm value of bacterial count (CFU) in the clay-amended phenanthrene and Cd-spiked soils over 21 days of incubation (letters above bars show the significant differences between adsorbents for each Cd level on each sampling day; ns = not significant).

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Fig. 5. Soil respiration (CO2 evolution) pattern from the clay- and modified clay-amended mixed-contaminated soils over 21 days of incubation.

tamination reduced bacterial growth in soils by 24% without clay treatment and 4% with bentonite treatment. In contrast, AB and ABP applications overcame the Cd inhibition; they improved the microbial proliferation by 22% and 10%, respectively, at a similar Cd level (Fig. 4). Despite ABP not being as good as AB in inducing bacterial growth at day 1, the former became a better performer from the 13th day onward. On day 13, ABP at Cd level 20 mg kg−1 showed the highest performance amounting to approximately 43% greater bacterial proliferation than the no-Cd treatment. At this stage, ABP showed 9% and 15% greater bacterial proliferation than B and AB treatments, respectively (p < 0.05). In ABP, the carboxylate groups hosted on the MIOC would have formed ligand complexes with Cd2+ and reduced metal toxicity to the soil bacteria. Furthermore, compared to the other adsorbents, ABP-treated soils contributed larger amounts of TOC (Table 1) due to a greater organic loading in the material, which provided a compatible growth environment for microorganisms. In general, CFU gradually increased as the incubation period progressed. However, after 21 days, CFU counts were slightly lower in AB- and ABP-treated soils than B- or NC-treated soils. The gradual depletion of the preferred carbon source (phenanthrene) to the soil bacteria and/or any microbial community shift in the heterogeneous system due to the change in the bioavailability of carbon source could effect this change [39]. 3.2.2. Soil respiration Like CFU (Figs. 3 and 4) soil respiration was also improved due to modified-clay applications to the mixed-contaminated soils (Fig. 5). Soil respiration is contributed by all soil organisms including plant roots present in a unit mass of soil at specific moisture content per unit time [40]. All treatments in this study were maintained uniformly at 60% WHC of the clay-amended soils and any CO2 -efflux contributed by plant roots was carefully excluded during soil preparation. Irrespective of Cd contamination levels in the mixed-contaminated soils, the ABP-treated soil maintained an elevated CO2 -efflux (soil respiration) compared to B- and ABtreated soils throughout the incubation period (Fig. 5). Since ABP could limit Cd bioavailability through chelation (Section 3.3), it promoted microbial proliferation and activity. Initially respiration rates in all the treatments were higher, which gradually declined with the incubation time (Fig. 5). The C-mineralisation pattern in a contaminated-soil largely depends on the abundance and

type of carbon sources. In this study, easily degradable native or introduced carbon sources, including phenanthrene, were likely to become exhausted toward the latter part of the incubation period. By employing microbial consortium in a PAH-spiked soil, Silva et al. [41] observed a similar CO2 -efflux pattern which indicated a strong correlation between CO2 evolution and PAH bioremediation. 3.2.3. Dehydrogenase activity (DHA) Microbial degradation of PAHs in soil is mediated by various enzymatic functions which are governed mainly by bacteria and fungi [42]. Soil DHA is considered as one of the most important indicators of microbial population. The DHA data in this study further revealed that ABP when applied to the mixed-contaminated soils facilitated microbial activity better than the control and other adsorbents (Fig. 6). DHA in the field soil (2.42 ± 0.08 ␮g TPF g−1 dry soil h−1 ) was reduced by about 45% due to phenanthrene spiking (prior to incubation). In ABP-amended soil, it was reduced up to 41%, which was less than in the other two treatments (about 55% and 63% for B and AB, respectively). These differences were statistically insignificant (p > 0.05). However, during incubation, DHA showed a variable trend as a result of the various amendments. Among all the treatments, only ABP-amendment enhanced the TPF values at days 1, 13 and 21 (2.14 ± 0.09, 0.90 ± 0.04 and 1.09 ± 0.08 ␮g g−1 dry soil h−1 , respectively) in the mixedcontaminated soil having the highest Cd loading (20 mg kg−1 ). These values were significantly greater (p < 0.05) than those of NC-, B- and AB-treated soils. The overall interactive effect (based on a univariate analysis) of different adsorbents in soils containing different Cd levels was not significant over the entire incubation period (p > 0.05). The main effect of either adsorbent or Cd concentration on DHA was also insignificant (p > 0.05) over this period. However, the adsorbents showed significant differences (p < 0.05) at each Cd concentration at a given sampling date (days 1, 7, 13 and 21) (Fig. 6). This indicated that different clay treatments could provide different environments to soil microorganisms at different Cd levels. The bacterial CFU, however, insignificantly correlated with the corresponding DHA value due to a possible discrepancy between the cultivable bacteria and total number of soil microorganisms [43]. DHA is an intracellular enzyme which takes part in the oxidative metabolism of soil microorganisms. In addition to bacteria, soil fungi, which are another important group of PAH-degraders [44], may contribute to

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Fig. 6. Dehydrogenase activity in clay- and modified clay-amended mixed-contaminated soils over 21 days of incubation (letters above bars show the significant differences between adsorbents for each Cd level on each sampling day).

the overall DHA. Nevertheless, like CFU and soil respiration, DHA was also enhanced by the ABP-amendment through safeguarding the microorganisms from elevated Cd toxicity. 3.3. Bioavailable Cd The bioavailable Cd fraction was significantly lower in the MIOC-treated soil than the control and other clay-amended soils (Fig. 7). At day 1, ABP demonstrated the maximum Cd immobilization (p < 0.05) at 10 mg kg−1 Cd level leaving only 18.16 ± 1.56% Cd bioavailable, whereas, NC, B and AB treatments showed 33.46 ± 4.34, 36.12 ± 6.45 and 33.09 ± 1.58% bioavailable Cd, respectively. The highest Cd retention by ABP could enhance the CFU and DHA during the given incubation time (Figs. 4 and 6). This retention capacity declined slowly over time. However, at both Cd concentrations (10 and 20 mg kg−1 ), its bioavailability remained lower due to ABP application, which was significant on day 13 at 20 mg kg−1 Cd level (p < 0.05) (Fig. 7). The higher CFU value on that day (Fig. 4) at the same Cd level perfectly synchronized with lower Cd bioavailability in the soil. At completion of incubation (day 21), the relationship between Cd bioavailability and CFU/DHA/respiration was, however, inconsistent (Figs. 4–7) due to a possible microbial community shift and/or exhaustion of food, which warrants further research [39,45]. However, this study clearly indicated that the MIOC could reduce Cd toxicity in a short reaction time and thus enhance microbial growth and activity in mixed-contaminated soils as seen immediately after incubation at day 1 (Fig. 8). Should the MIOC simultaneously maintain phenanthrene in bioavailable form, it would facilitate/enhance complete

Fig. 7. Bioavailable fraction of Cd in clay- and modified clay-amended soils in microcosm over 21 days of incubation (letters above bars show the significant differences between adsorbents for each Cd level on each sampling day; ns = not significant)

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Fig. 8. The microbial response (CFU and dehydrogenase activity) to the bioavailable Cd in clay and modified clay-amended soil for each Cd level at day 1.

Fig. 9. Bioavailable fraction of phenanthrene in clay- and modified clay-amended sterile soils in microcosm over 21 days of incubation (letters above bars show the significant differences between adsorbents for each Cd level on each sampling day).

removal of this contaminant by the action of highly proliferated native and/or introduced microorganisms. 3.4. Bioavailable phenanthrene The bioavailability of phenanthrene in the sterile soil was significantly higher (p < 0.05) in ABP-treatment than in AB-, B- and NC-treatments over the entire incubation period (Fig. 9). Afterwards the bioavailable fraction in all the treatments reduced gradually due to possible aging effect (no opportunity for degradation here since a sterile autoclaved soil was used) [46]. However, compared to other treatments (NC, B and AB), ABP maintained a significantly greater phenanthrene bioavailability in the soils (Fig. 10). At completion of the incubation (day 21), ABP-amended soil yielded 55% bioavailable phenanthrene, while AB-, B- and NCtreated soils yielded 41%, 3% and 2%, respectively. This clearly supports the hypothesis that the MIOC (ABP) maintains phenanthrene in bioavailable form along with reducing Cd-toxicity. Through a metal complexion mechanism, ABP appeared to selectively immobilize Cd in a Cd-phenanthrene mixedcontaminated soil [17] giving less room for phenanthrene adsorption. The bioavailable phenanthrene supported microbial growth by supplying food to the potential degraders (Fig. 3 and 4). Interestingly, the depletion of available phenanthrene (about 13–17%) in the microbially-active soil in comparison to the sterile soil at day 21 appeared significant in ABP treatment (p < 0.05) (Fig. 10). This depletion, which was calculated following subtracting the effect of ageing (as in Fig. 9), was due to the mineralization of phenanthrene because the sterile soil had no active microorganism. The C:N ratio of ABP-amended soils measured on day 21 showed

Fig. 10. Bioavailable fraction of phenanthrene in clay- and modified clay-amended soils in microcosm at day 21 in reference to sterile soil (letters above bars show the significant differences between sterile and treated soils for each adsorbent; ns = not significant).

slightly higher values than other treatments (Table 1), which also indicated a PAH mineralization process favorably occurring in these soils [47]. 4. Conclusions The organic modification of bentonite with Arquad® 2HT-75 followed by PA changed its structure and surface properties, which brought about a net increase in negative surface charge in the MIOC. This negative charge over a wide environmental pH range (pH 2–10) enabled the MIOC to adsorb/immobilize heavy metal (Cd) in a mixed-contaminated soil. This also enhanced the microbial activities in the mixed-contaminated soil by limiting metal bioavailability in the solution phase (thereby reducing metal toxicity to microorganisms) and facilitating the supply of suitable substrate (enhancing phenanthrene bioavailability and mineralization) for microbial proliferation. Since such a MIOC has huge potential in achieving efficient PAH biodegradation in metal cocontaminated soils, future research needs to address: (a) the compatibility of MIOC with biological functions under complex bio/physico-chemical environment in historically contaminated soils,

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(b) the stability and durability of the MIOC in immobilizing metal cations, and (c) the microbial community shift and changes in the function of housekeeping genes as a result of MIOC application to a mixed-contaminated soil. Acknowledgements This study was conducted as a part of BB’s PhD research under the supervision of BS and RN. BB is thankful to the University of South Australia for awarding him an International President’s Scholarship. The authors also acknowledge CRC CARE for providing infrastructural support. AM is grateful to Dr Ashok Patra (Director) and Dr A Subba Rao (Ex-Director), Indian Institute of Soil Science, and acknowledges ICAR for awarding him a NAIP international training grant. References [1] M. Hoch, Organotin compounds in the environment – an overview, Appl. Geochem. 16 (2001) 719–743. [2] B. Ramakrishnan, M. Megharaj, K. Venkateswarlu, N. Sethunathan, R. Naidu, Mixtures of environmental pollutants: effects on microorganisms and their activities in soils, in: D.M. Whitacre (Ed.), Reviews of Environmental. Contamination and Toxicology, Springer, New York, 2011, pp. 63–120. [3] P. Thavamani, M. Megharaj, R. Naidu, Multivariate analysis of mixed contaminants (PAHs and heavy metals) at manufactured gas plant site soils, Environ. Monit. Assess. 184 (2012) 3875–3885. [4] N.S. Bolan, D.C. Adriano, R. Naidu, Role of phosphorus in (im) mobilization and bioavailability of heavy metals in the soil-plant system, Rev. Environ. Contam. Toxicol. 177 (2003) 1–44. [5] J.C. Philp, S.M. Bamforth, I. Singleton, R.M. Atlas, Environmental pollution and restoration: a role for bioremediation, in: R.M. Atlas, J.C. Philp (Eds.), Bioremediation: Applied Microbial Solutions for Real-World Environmental Cleanup, ASM press, Washington, D.C, 2005, pp. 1–48. [6] K. Vig, M. Megharaj, N. Sethunathan, R. Naidu, Bioavailability and toxicity of cadmium to microorganisms and their activities in soil: a review, Adv. Environ. Res. 8 (2003) 121–135. [7] M. Lu, K. Xu, J. Chen, Effect of pyrene and cadmium on microbial activity and community structure in soil, Chemosphere 91 (2013) 491–497. [8] T.R. Sandrin, R.M. Maier, Impact of metals on the biodegradation of organic pollutants, Environ. Health Perspect. 111 (2003) 1093–1101. [9] B. Sarkar, Y. Xi, M. Megharaj, G.S.R. Krishnamurti, D. Rajarathnam, R. Naidu, Remediation of hexavalent chromium through adsorption by bentonite based Arquad® 2HT-75 organoclays, J. Hazard. Mater. 183 (2010) 87–97. [10] V.K. Gupta, A. Nayak, B. Bhushan, S. Agarwal, A critical analysis on the efficiency of activated carbons from low cost precursors for heavy metals remediation, Crit. Rev. Environ. Sci. Technol. (2014) http://dx.doi.org/10.1080/ 10643389.2013.876526 [11] D. Mohan, H. Kumar, A. Sarswat, M. Alexandre-Franco, C.U. Pittman Jr., Cadmium and lead remediation using magnetic oak wood and oak bark fast pyrolysis bio-chars, Chem. Eng. J. 236 (2014) 513–528. [12] B. Sarkar, Y.F. Xi, M. Megharaj, G.S.R. Krishnamurti, M. Bowman, H. Rose, R. Naidu, Bioreactive organoclay: a new technology for environmental remediation, Crit. Rev. Environ. Sci. Technol. 42 (2012) 435–488. [13] A.R.A. Usman, Y.A. Almaroai, M. Ahmad, M. Vithanage, Y.S. Ok, Toxicity of synthetic chelators and metal availability in poultry manure amended Cd, Pb and As contaminated agricultural soil, J. Hazard. Mater. 262 (2013) 1022–1030. [14] G. Mühlbachová, Soil microbial activities and heavy metal mobility in long-term contaminated soils after addition of EDTA and EDDS, Ecol. Eng. 37 (2011) 1064–1071. [15] J.M. Thomas, M. Alexander, Colonization and mineralization of palmitic acid by Pseudomonas pseudoflava, Microb. Ecol. 14 (1987) 75–80. [16] N. Bolan, A. Kunhikrishnan, R. Thangarajan, J. Kumpiene, J. Park, T. Makino, M.B. Kirkham, K. Scheckel, Remediation of heavy metal(loid) s contaminated soils – to mobilize or to immobilize? J. Hazard. Mater. 266 (2014) 141–166. [17] P. Malakul, K.R. Srinivasan, H.Y. Wang, Metal adsorption and desorption characteristics of surfactant-modified clay complexes, Ind. Eng. Chem. Res. 37 (1998) 4296–4301. [18] B. Sarkar, M. Megharaj, D. Shanmuganathan, R. Naidu, Toxicity of organoclays to microbial processes and earthworm survival in soils, J. Hazard. Mater. 261 (2013) 793–800. [19] B. Sarkar, M. Megharaj, Y. Xi, G.S.R. Krishnamurti, R. Naidu, Sorption of quaternary ammonium compounds in soils: implications to the soil microbial activities, J. Hazard. Mater. 184 (2010) 448–456. [20] P. Malakul, K.R. Srinivasan, H.Y. Wang, Metal toxicity reduction in naphthalene biodegradation by use of metal-chelating adsorbents, Appl. Environ. Microbiol. 64 (1998) 4610–4613.

137

[21] J.D. Van Nostrand, T.J. Khijniak, B. Neely, M.A. Sattar, A.G. Sowder, G. Mills, P.M. Bertsch, P.J. Morris, Reduction of nickel and uranium toxicity and enhanced trichloroethylene degradation to Burkholderia vietnamiensis PR1301 with hydroxyapatite amendment, Environ. Sci. Technol. 41 (2007) 1877–1882. [22] M. Addy, B. Losey, R. Mohseni, E. Zlotnikov, A. Vasiliev, Adsorption of heavy metal ions on mesoporous silica-modified montmorillonite containing a grafted chelate ligand, Appl. Clay Sci. 59–60 (59) (2012) 115–120. [23] A.R. Johnsen, L.Y. Wick, H. Harms, Principles of microbial PAH-degradation in soil, Environ. Pollut. 133 (2005) 71–84. [24] B. Sarkar, M. Megharaj, Y. Xi, R. Naidu, Structural characterization of Arquad® 2HT-75 organobentonites: surface charge characteristics and environmental application, J. Hazard. Mater. 195 (2011) 155–161. [25] U.C. Brinch, F. Ekelund, C.S. Jacobsen, Method for spiking soil samples with organic compounds, Appl. Environ. Microbiol. 68 (2002) 1808–1816. [26] G.E. Rayment, D.J. Lyons, Soil Chemical Methods – Australasia, CSIRO Publishing, Collingwood, VIC, 2011. [27] S.K. Chaerun, K. Tazaki, R. Asada, K. Kogure, Interaction between clay minerals and hydrocarbon-utilizing indigenous microorganisms in high concentrations of heavy oil: implications for bioremediation, Clay Miner. 40 (2005) 105–114. [28] J. Bloem, D.W. Hopkins, A. Benedetti, Microbiological Methods for Assessing Soil Quality, CABI Publishing, Wallingford, Oxfordshire, UK, 2005. [29] A. Pascaud, M.L. Soulas, S. Amellal, G. Soulas, An integrated analytical approach for assessing the biological status of the soil microbial community, Eur. J. Soil Biol. 49 (2012) 98–106. [30] L.E. Casida, D.A. Klein, T. Santoro, Soil dehydrogenase activity, Soil Sci. 98 (1964) 371–376. [31] M.V. Ruby, R. Schoof, W. Brattin, M. Goldade, G. Post, M. Harnois, D.E. Mosby, S.W. Casteel, W. Berti, M. Carpenter, D. Edwards, D. Cragin, W. Chappell, Advances in evaluating the oral bioavailability of Inorganics in soil for use in human health risk assessment, Environ. Sci. Technol. 33 (1999) 3697–3705. [32] S. Brown, B. Christensen, E. Lombi, M. McLaughlin, S. McGrath, J. Colpaert, J. Vangronsveld, An inter-laboratory study to test the ability of amendments to reduce the availability of Cd, Pb, and Zn in situ, Environ. Pollut. 138 (2005) 34–45. [33] J.W. Kelsey, B.D. Kottler, M. Alexander, Selective chemical extractants to predict bioavailability of soil-aged organic chemicals, Environ. Sci. Technol. 31 (1996) 214–217. [34] L. Duan, T. Palanisami, Y. Liu, Z. Dong, M. Mallavarapu, T. Kuchel, K.T. Semple, R. Naidu, Effects of ageing and soil properties on the oral bioavailability of benzo[a]pyrene using a swine model, Environ. Int. 70 (2014) 192–202. [35] Y.H. Shin, S.M. Chou, C. Lin, Sorption mechanisms of selected chlorinated volatile organic compounds onto organoclays, J. Environ. Eng. Manage. 20 (2010) 161–168. [36] R. Naidu, N.S. Bolan, R.S. Kookana, K.G. Tiller, Ionic-strength and pH effects on the sorption of cadmium and the surface charge of soils, Eur. J. Soil Sci. 45 (1994) 419–429. [37] U.C. Ugochukwu, Biodegradation of crude oil hydrocarbons supported on clay minerals, in: Faculty of Science, Agriculture and Engineering, Newcastle University, Newcastle upon Tyne, UK, 2011. [38] L.E. McAllister, K.T. Semple, Role of clay and organic matter in the biodegradation of organics in soil, in: M. Mandl, L.L. Barton, A. Loy (Eds.), Geomicrobiology: Molecular and Environmental Perspective, Springer Netherlands, Dordrecht, 2010, pp. 367–384. ˜ [39] M. Vinas, J. Sabaté, M.J. Espuny, A.M. Solanas, Bacterial community dynamics and polycyclic aromatic hydrocarbon degradation during bioremediation of heavily creosote-contaminated soil, Appl. Environ. Microbiol. 71 (2005) 7008–7018. [40] L. Yiqi, X. Zhou, Soil Respiration and the Environment, Elsevier Science, Burlington, 2010. [41] I.S. Silva, E.d.C.d. Santos, C.R.d. Menezes, A.F.d. Faria, E. Franciscon, M. Grossman, L.R. Durrant, Bioremediation of a polyaromatic hydrocarbon contaminated soil by native soil microbiota and bioaugmentation with isolated microbial consortia, Bioresour. Technol. 100 (2009) 4669–4675. [42] A. Haritash, C. Kaushik, Biodegradation aspects of polycyclic aromatic hydrocarbons (PAHs): a review, J. Hazard. Mater. 169 (2009) 1–15. [43] P. Olga, K. Petar, M. Jelena, R. Srdjan, Screening method for detection of hydrocarbon-oxidizing bacteria in oil-contaminated water and soil specimens, J. Microbiol. Methods 74 (2008) 110–113. [44] I.S. Morelli, M.C.N. Saparrat, M.T. Del Panno, B.M. Coppotelli, A. Arrambari, Bioremediation of PAH-contaminated soil by fungi, in: Fungi as Bioremediators, Springer, 2013, 2015, pp. 159–179. [45] C.H. Nakatsu, N. Carmosini, B. Baldwin, F. Beasley, P. Kourtev, A. Konopka, Soil microbial community responses to additions of organic carbon substrates and heavy metals (Pb and Cr), Appl. Environ. Microbiol. 71 (2005) 7679–7689. [46] B.J. Reid, K.C. Jones, K.T. Semple, Bioavailability of persistent organic pollutants in soils and sediments – a perspective on mechanisms, consequences and assessment, Environ. Pollut. 108 (2000) 103–112. [47] N. Leys, L. Bastiaens, W. Verstraete, D. Springael, Influence of the carbon/nitrogen/phosphorus ratio on polycyclic aromatic hydrocarbon degradation by Mycobacterium and Sphingomonas in soil, Appl. Microbiol. Biotechnol. 66 (2005) 726–736.