Environ Sci Pollut Res (2013) 20:22–30 DOI 10.1007/s11356-012-1016-8

USE OF IRON AND OTHER TRANSITION METALS IN ENVIRONMENTAL REMEDIATION PROCESSES

Treatment of hydrocarbon contamination under flow through conditions by using magnetite catalyzed chemical oxidation M. Usman & P. Faure & C. Lorgeoux & C. Ruby & K. Hanna

Received: 15 March 2012 / Accepted: 25 May 2012 / Published online: 9 June 2012 # Springer-Verlag 2012

Abstract Soil pollution by hydrocarbons (aromatic and aliphatic hydrocarbons) is a major environmental issue. Various treatments have been used to remove them from contaminated soils. In our previous studies, the ability of magnetite has been successfully explored to catalyze chemical oxidation for hydrocarbon remediation in batch slurry system. In the present laboratory study, column experiments were performed to evaluate the efficiency of magnetite catalyzed Fenton-like (FL) and activated persulfate (AP) oxidation for hydrocarbon degradation. Flow-through column experiments are intended to Responsible editor: Philippe Garrigues M. Usman (*) Institute of Soil and Environmental Sciences, University of Agriculture, Faisalabad 38040, Pakistan e-mail: [email protected] M. Usman : C. Ruby : K. Hanna Laboratoire de Chimie Physique et Microbiologie pour l’Environnement, LCPME, UMR 7564 CNRS–Université de Lorraine, 405 rue de Vandoeuvre, 54600 Villers Les Nancy, France M. Usman : P. Faure : C. Lorgeoux Géologie et Gestion des Ressources Minérales et Energétiques, G2R, UMR 7566, CNRS–Université de Lorraine, 54506 Vandoeuvre Les Nancy, France K. Hanna Ecole Nationale Supérieure de Chimie de Rennes, UMR CNRS 6226 “Sciences Chimiques de Rennes”, Avenue du Général Leclerc, 35708 Rennes Cedex 7, France K. Hanna Université Européenne de Bretagne, Rennes, France

provide a better representation of field conditions. Organic extracts isolated from three different soils (an oilcontaminated soil from petrochemical industrial site and two soils polluted by polycyclic aromatic hydrocarbon (PAH) originating from coking plant sites) were spiked on sand. After solvent evaporation, spiked sand was packed in column and was subjected to oxidation using magnetite as catalyst. Oxidant solution was injected at a flow rate of 0.1 mL min−1 under water-saturated conditions. Organic analyses were performed by GC–mass spectrometry, GC–flame ionization detector, and microFourier transform infrared spectroscopy. Significant abatement of both types of hydrocarbons (60–70 %) was achieved after chemical oxidation (FL and AP) of organic extracts. No significant by-products were formed during oxidation experiment, underscoring the complete degradation of hydrocarbons. No selective degradation was observed for FL with almost similar efficiency towards all hydrocarbons. However, AP showed less reactivity towards higher molecular weight PAHs and aromatic oxygenated compounds. Results of this study demonstrated that magnetite-catalyzed chemical oxidation can effectively degrade both aromatic and aliphatic hydrocarbons (enhanced available contaminants) under flowthrough conditions. Keywords Soil . Aliphatic hydrocarbons . Aromatic hydrocarbons . Magnetite . Oxidation . Fenton . Persulfate

Introduction The intense use of fossil organic matter (petroleum and coal), since eighteenth century for industrial purposes (petroleum extraction, refinery, coking plant, steel

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industries, etc.), has resulted in widespread pollution by polycyclic aromatic hydrocarbons (PAHs) and aliphatic hydrocarbons (oil by-products) that cause environmental and health concerns. Sixteen PAHs (US Environmental Protection Agency (US EPA) list), in particularly, are considered as priority pollutants by US EPA and European community and are well-known carcinogenic pollutants (ITRC 2005). Different remediation techniques have been tested for the removal of PAHs from contaminated matrices. Due to their high persistence in soil, they are resistant to environmental degradation and imply huge industrial treatments. On the contrary, aliphatic hydrocarbons are easily degradable by bioremediation (Prince 1993). However, complete mineralization of oil to CO2 and H2O cannot be achieved by bioremediation and always leaves more or less complex residues (Atlas 1995). Moreover, biorefractory materials, especially asphaltenes, were recalcitrant to biodegradation (Gough and Rowland 1990; Chaillan et al. 2006). The widespread distribution of such recalcitrant aliphatic hydrocarbons and PAHs has resulted in increasing interest to develop methods to degrade hydrocarbons in contaminated sites. Application of chemical oxidation methods, especially Fenton and persulfate treatments, has been quite successful in remediating hydrocarbon contamination (Watts and Dilly 1996; Kong et al. 1998; Watts et al. 2002; Kanel et al. 2004; Rivas 2006; Ferrarese et al. 2008; Do et al. 2010; Yen et al. 2011). Both of these oxidants require activation by iron (soluble FeII or iron minerals) in order to produce stronger radicals. As soluble FeII could become inactive as a result of its rapid oxidation and precipitation, chelating agents were used to keep this FeII in solution (Liang et al. 2004; Ferrarese et al. 2008). Use of iron minerals also provided another alternative to produce stronger radicals at circumneutral pH (Kong et al. 1998; Kanel et al. 2004; Matta et al. 2007; Hanna et al. 2008; Do et al. 2010). Recently, highest catalytic reactivity was shown by FeII-bearing minerals like magnetite (Fe3O4) as compared to the only FeIII oxides for heterogeneous catalytic oxidation of organic pollutants (Kong et al. 1998; Matta et al. 2007; Hanna et al. 2008; Xue et al. 2009a). It also showed strong structural and catalytic stability and, therefore, can be used for further reaction cycle (Xue et al. 2009a). In our batch studies, magnetite (FeII–Fe2IIIO4) was found effective to catalyze chemical oxidation for significant degradation of aliphatic hydrocarbons (Usman et al. 2012a) and PAHs (Usman et al. 2012b, c) at circumneutral pH. However magnetite-catalyzed Fenton and persulfate oxidation has never been applied for hydrocarbon degradation under flow-through conditions. Tests under flow-through conditions when oxidant is

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injected downwards or upwards are closer to fieldscale application (e.g., in situ treatments ISCO). Therefore, in the present study, column experiments were conducted to evaluate the efficiency of chemical oxidation to degrade hydrocarbons. As soluble FeII was found unable to act as a catalyst for Fenton reaction as well as persulfate oxidation at circumneutral pH (Usman et al. 2012a, b, c), only the use of magnetite as a catalyst was evaluated in column experiments. Moreover, our previous findings (bench scale) showed that no PAH oxidation can be achieved w ithout enhancing PAH availability through solvent extraction (Usman et al. 2012a, b). For this reason, this study comprises of testing the catalytic ability of magnetite for the degradation of hydrocarbons in organic extracts under flowthrough conditions. Organic extracts were isolated from three soils (one oil-contaminated soil and two different PAH-contaminated soils), and each was spiked on sand used as inert support. Oxidation was performed by injecting oxidants (H2O2 and sodium persulfate) at fixed flow rate (0.1 mL min−1) in sand-packed column containing magnetite. For column tests, the hydrodynamic parameters such as flow rate, pore velocity, and bed height were chosen to ensure a convective regime. Degradation of hydrocarbons was monitored by GC–mass spectrometry (GC-MS), GC–flame ionization detector (GC-FID), and micro-Fourier transform infrared spectroscopy (μ-FTIR).

Experimental section Sample preparation Magnetite-rich sand (MRS) (10 % of magnetite w/w) used was prepared and characterized in context of a previous study (Usman et al. 2012a). Oxidative degradation of aliphatic hydrocarbons and PAHs was studied in organic extracts from hydrocarbon-contaminated soils. Aliphatic hydrocarbons present in weathered oil were extracted from an oil-contaminated forest soil from Pechelbronn oil field (located in Alsace, France) that was under natural attenuation. Organic extracts containing PAHs were isolated from two polluted soils of former coking plant sites (Homécourt (H) and NeuvesMaisons (NM) located in the northeast of France). For organic extraction, contaminated soil samples were sieved at 2 mm, freeze dried, ground to 500 μm, and extracted using an automatic extractor Dionex® ASE 200 (accelerated solvent extractor) at 100 °C and 130 bars with dichloromethane (DCM). Each organic extract (WO, H, and NM samples) was spiked on MRS to obtain a final concentration of 4 gkg−1

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Oxidation under flow-through conditions Three treatments were included to study oxidative degradation of hydrocarbons which were: (1) Fenton-like (H2O2 + magnetite 0 FL), (2) activated persulfate (sodium persulfate + magnetite 0 AP), and (3) blank experiments by using magnetite alone without any oxidant. The schematic diagram of column used in this study is given in Fig. 1. In a glass chromatographic column of 40 cm length and 2.6 cm internal diameter (XK 26/20, GE Healthcare), the spiked MRS particles were packed to a height of 6.5 cm, corresponding to a dry mass of 50 g. The dry porous bed had a uniform bulk density (ρ) of 1.23±0.01 g/cm3. After packing, the column was cautiously wetted upward with the Milli-Q water (Millipore). Throughout the experiments, the flow rate was held constant at 0.1 mL min−1, corresponding to a pore water velocity of 0.019 cm min−1; the flow direction was from the bottom to top of the column in order to ensure water-saturated conditions. Blank experiments (prepared in similar way) were conducted by injecting water in sand-packed column. The outflow concentration of dissolved organic carbon (DOC) was found almost zero in its outflow, stating the absence of hydrocarbons mobilization from column. Flow rate was checked several times by comparing inflow and outflow and was found constant. Porous volume of the MRS column measured by weighing before and after water saturation was almost 10.5±0.02 mL. The flow characteristics of the porous

m/m0

of MRS. The DCM was evaporated with a continuous mixing to ensure homogeneous contaminant distribution. This spiked sand was considered as reference (t00).

Fig. 2 Extractable organic matter content of organic extracts of contaminated soils (WO, H, and NM) before (empty square, t00) and after oxidation by H2O2 + magnetite (black-filled square, FL) and sodium persulfate+magnetite (gray-filled square, AP). These organic extracts were spiked on sand and were subjected to oxidation under flowthrough conditions. Blank experiment was performed in the presence of magnetite, but without any oxidant. This degradation is represented in terms of m/m0 where m is the amount of EOM after oxidation, and m0 is EOM at t00 (before oxidation)

bed determined by a nonreactive tracer experiment (Br) were found homogeneous without preferential path. This is in agreement with previous studies where similar porous media was tested (Hanna et al. 2010). Dispersivity was found around 150 μm close to particle size of used sand. Peclet number was found higher than 300, indicating the predominance of a convective regime. DOC was almost zero in saturation experiments stating the absence of carbon mobilization. To conduct oxidation, oxidant solution (total injected volume0500 mL) with concentration equivalent to oxidant/Fe molar ratio equal to 10:1 and 1:1 for FL and AP, respectively, was injected with same flow rate (0.1 mL min−1) under water-saturated conditions. These oxidant/Fe ratios were selected according to previously published works (Liang et al. 2004; Xue et al. 2009a; Usman et al. 2012a, b, c). In all experiments, dissolved iron concentrations in the outflow were measured by

Spiked sand

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Peristaltic pump FL oxidation AP oxidation

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Fig. 1 Diagram showing the layout of column used for oxidation experiments. Oxidants were injected in column packed with spiked magnetite rich sand

Fig. 3 GC-FID chromatograms of weathered oil before (t00) and after oxidation by H2O2 + magnetite (FL) and sodium persulfate + magnetite (AP). Intensities of chromatograms are proportional to the oil concentration

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inductively coupled plasma atomic emission spectroscopy. Outflow concentration of iron was negligible, indicating that no magnetite had been flushed out from the column. Once whole solution was injected, and the solid samples were freeze dried to remove water. Three replicates were analyzed from the same column. All results were expressed as a mean value of the three values and standard deviation of the three replicates was less than 5 %. All experiments were performed without pH adjustment (6.8±0.3) and at room temperature (20–25 °C). Extraction and analysis After freeze drying, the samples were submitted to five cycles of CHCl3 Soxhlet extractions for 24 h each. The solvent volume was then reduced to 50 mL under a nitrogen flow, and then 5 mL of the solution was dried and weighed in order to determine the amount of extractable organic matter (EOM). The hydrocarbon oil index (HI) was measured for oil hydrocarbons according to ISO 16703:2004 procedure using a GC–FID 7890 Agilent Technologies. GC-MS quantification of PAHs in organic extracts was performed by adding internal standards to the samples. An internal deuterated PAHs standard mix (naphthalene-D8, acenaphthene-D10, phenanthrene-D10, chrysene-D12, and perylene-D12, supplied by Cluzeau) was added. A 1 μL

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Results and discussion Degradation of oil hydrocarbons Oxidation of hydrocarbons was conducted by applying H2O2 + magnetite (FL) and sodium persulfate + magnetite (AP). Effluents from columns after oxidation were collected and analyzed for DOC, but its value was very LMW-PAHs

HMW-PAHs

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PAH contents (µg.g -1 dw)

Fig. 4 Contents of individual PAHs (US EPA list) in H and NM organic extracts (spiked on sand) at t00 before oxidation (empty square) and after oxidation by FL (H2O2 + magnetite + black-filled square) and AP oxidation (sodium persulfate + magnetite 0 grayfilled square). PAH contents are based on the measurements by GC-MS

amount of solution was then injected into an Agilent Technologies 6890 gas chromatograph equipped with a DB-5 MS (length 60 m, diameter 0.25 mm, film thickness 0.1 μm) capillary column coupled to an Agilent Technologies 5973 mass spectrometer operating in full scan mode. The temperature program was the following: 70 to 250 °C at 15 °C min−1, then 250 to 315 °C at 3 °C min−1, and 60 min holds at 315 °C. The carrier gas was helium at 1.5 mL min−1 constant flow. The micro-Fourier Transform Infrared spectroscopic analysis were performed on an infrared spectrometer Bruker IFS55 coupled with a multipurpose Bruker IR microscope equipped with a MCT detector cooled with liquid N2. EOM were analyzed as described by Faure et al. (1999), using a diamond window. The spectra were recorded with the following conditions: size of the analyzed area 60 μm2, 64 accumulations (32 s), spectral resolution 4 cm−1, and gain 4.

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low (10–12 mg L−1), which represents less than 5 % of total organic matter. Negligible difference of DOC was observed between the different fractions of effluent and the different treatments (FL or AP). Therefore, degradation of oil hydrocarbons in WO was monitored by evolution of EOM and HI. The EOM evolution is represented by m/m0 before and after oxidation where m is the mass of organic extract after oxidation while m0 is the mass at t00 (Fig. 2). The EOM recovered at t00 was similar to initially added amount (~4 mg g−1), representing that no EOM was trapped in MRS as observed elsewhere (Usman et al. 2012a). Negligible decrease was observed in blank experiments (<3 %), while significant degradation (60–70 %) was obtained by FL and AP (Fig. 2). The HI was measured by GC– FID before and after oxidation, and results are in agreement with EOM observations (Fig. 3). Significant decrease in HI was obtained by FL (~73 %) and AP (~65 %), while almost no degradation in blank experiments was noticed. Results obtained in column

Fig. 5 FTIR spectra of organic extracts of soils contaminated by weathered oil and PAHs (H and NM soils) before and after oxidation by Fenton-like and activated persulfate

experiments confirm previous batch findings about the strong reactivity of magnetite to catalyze chemical oxidation (Usman et al. 2012a). However, slight decrease in degradation extent was observed as compared to batch slurry system (~84 % by FL and ~73 % by AP). This decrease in abatement extent is expected due to the kinetic limitation and restricted contact mode in column. Analyses by GC–FID revealed the existence of a large area of the raised baseline hump in sample chromatogram that describes an intense unresolved complex mixture (UCM) before oxidation (Fig. 3). It is generally considered as an iso and cycloalkanes mixture not resolved by chromatography (Gough and Rowland 1990). UCM was the sole reactant in the reference sample because of its totally refractory nature to microbial attack (Gough and Rowland 1990; Chaillan et al. 2006). Sampled soil area was under natural attenuation for years that resulted in the removal of n-alkanes that are easily biodegradable (Chaillan et al. 2006). These results reveal a significant efficiency of magnetite-

ν as CH 2 ν as CH 3

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both soils (H and NM) and were spiked on sand containing magnetite. Oxidation was initiated by injecting oxidant solutions (H2O2 and sodium persulfate) with flow rate of 0.1 mL min−1 in column experiments. The values of DOC in outflow of these columns were very low (10– 12 mg L−1), which is similar to that observed for WO oxidation. Before oxidation, the total content of PAHs in H extract was lower (~204 μg g−1) than NM extract with ~348 μg g−1 of sand (Fig. 4). The ratio of low molecular weight (LMW) PAH (sum of naphthalene to pyrene concentrations) over high molecular weight (HMW) PAH (sum of benzo[a]anthracene to benzo[g,h,i]perylene concentrations), LMW/HMW, was calculated. This ratio suggested that the composition of PAHs was different in both soils as represented by predominance of LMW-PAHs with a value of 3.81 in H extract as compared to 0.94 in NM extract with higher proportion of HMW-PAHs. Degradation of PAHs was monitored by evolution of EOM (Fig. 2) and content of 16 PAHs (Fig. 4). Blank experiments did not show degradation (<3 %), while oxidation treatments resulted in an intense decrease in EOM after oxidation (Fig. 2). Residual

catalyzed chemical oxidation to degrade refractory oil residues under flow-through conditions. Degradation of PAHs

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Fig. 6 Molecular distribution obtained by GC-MS of H and NM organic extract (spiked on sand) before oxidation (t00) and after oxidation by FL (H2O2 + magnetite) and AP oxidation (sodium persulfate+ magnetite). The compounds are detailed in Table 1

PHE

Oxidation treatments were conducted in lab-scale columns for the degradation of PAHs from soils of two ancient coking plant sites: H and NM. We already observed that no PAH degradation was found when both of these aged contaminated soils were subjected to oxidation (H2O2 or persulfate) whatever the catalyst used (soluble FeII or magnetite) (Usman et al. 2012b, c). It was found that PAH availability is the main limitation for this absence of degradation. This issue was addressed by pretreating these soils by a PAH availability enhancement agent (organic solvent or cyclodextrin). The application of an extraction pretreatment was found successful as a preliminary step before designing an oxidation treatment for an efficient PAH degradation in batch slurry system (Usman et al. 2012b, c). Therefore, organic extracts were isolated of

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concentrations of 16 PAHs achieved after oxidation treatments is represented in Fig. 4. As can be seen from data shown, both FL and AP were almost equally efficient for PAH degradation in H extract (~70 %). But in NM extract, FL showed higher degradation (~78 %) than AP (~63 %). This difference of degradation is due to the selective degradation of PAHs by AP with less efficiency towards HMW-PAHs. This selective behavior of AP resulted in less PAH degradation in NM extract (~63 %), which contains higher proportion of HMWPAHs. On the other hand, FL showed nonselective degradation without any preferential behavior for LMW- or HMW-PAHs although HMW-PAHs are less susceptible to degradation due to their more hydrophobic nature and stronger sorption properties onto micropores of particulates in soil matrices (Rivas 2006; Gan et al. 2009). This difference in reactivity of both oxidants (FL and AP) could be attributed to the possible production of surfactants during Fenton oxidation of PAHs. Indeed, partial oxidation of hydrocarbons and/or native organics result in the formation of compounds having surfactant-like properties (Ndjou'ou and Cassidy 2006; Gryzenia et al. 2009). These surfactants could solubilize HMW-PAHs to render them chemically more available for their better removal (Gryzenia et al. 2009). Moreover, reductants (e.g., superoxide) can be formed during Fenton oxidation which could act as surfactant and contribute in enhanced desorption of PAHs (Watts et al. 1999). Thus, all these factors probably result in nonselective behavior of FL to degrade PAHs. In tested soils, nonselective PAH degradation was observed during air oxidation of PAHs (Biache et al. 2011), while thermal desorption treatment showed less degradation of HMW-PAHs (Biache et al. 2008). To our best knowledge, no data about PAH oxidation (neither in organic extract nor in soil) in column is available in the literature. This makes difficult to compare the findings of this study with the literature. However, these results consistently show that both oxidants, when catalyzed by magnetite, could result in significant degradation of PAHs (enhanced available) under both batch and flow-through conditions. Magnetite has exhibited excellent structural and catalytic stabilities, and therefore, it can be used for several oxidation cycles (Xue et al. 2009a, b). Investigation of possible by-products For all oxidation experiments, formation of by-products was investigated by μ-FTIR (Fig. 5) and GC-MS (Fig. 6). Investigations by μ-FTIR (Fig. 5) revealed that aliphatic bands are dominated in WO organic extract (ν C H a l i 3, 00 0– 2, 80 0 c m − 1 an d δ C H a l i 1 ,4 70 – 1,360 cm−1) with a νCH3/νCH2 ratio values compatible with the predominance of iso- and cycloalkanes

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corresponding to an unresolved complex mixture already observed by GC-FID (Fig. 3). Even if one weak oxygenated band (νC 0 O, 1,745–1,705 cm−1) appeared in all samples, initial EOM characteristics remain similar after treatment by both oxidants, stating the absence of significant by-products. Initial EOM spectra of both H and NM organic extracts (Fig. 5) are characterized by intense aromatic bands (νCHaro, 3,100–3,000 cm−1; νC 0 C, 1,620– 1,590 cm−1; and γCHaro, 900–700 cm−1) and in lesser extent aliphatic (νCHali 3,000–2,800 cm−1 and δCHali 1,470–1,360 cm−1) and oxygenated bands (especially νOH, 3,700–3,100 cm − 1 and νC 0 O, 1,745– 1,705 cm−1). Moreover, H organic extract exhibits a specific spectral signature due to the occurrence of nitrogen-bearing bands (νN–H 3,700–3,600 cm−1 and νC–N 1,300–1,100 cm−1). Moreover, oxygenated (dibenzofuran (DBF) and fluorenone) and substituted compounds (e.g., C1-N, C2-N, methyldibenzofuran (C1DBF), and C1-PHE) were observed along with 16 PAHs in GC-MS chromatograms (Fig. 6). Except the disappearance of nitrogen-bearing bands in the case of H organic extract, initial EOM characteristics remain Table 1 Details of the abbreviated compounds enlisted in the chromatograms of Fig. 6 Code

Name

IS1 C1-N C2-N IS2 ACP DBF C1-DBF FLO DBT IS3 PHE ANT ANTO C1-PHE

Naphthalene-D8 Methyl-naphthalene Ethyl and dimethylnaphthalene Acenaphtene-D10 Acenaphtene Dibenzofuran Methyldibenzofuran Fluorenone Dibenzothiopene Phenanthrene-D10 Phenanthrene Anthracene Anthracenone Methylphenanthren

ANTDO FLUO PYR BNF BAA IS4 CHY 252 IS5 276

Anthracendione Fluoranthene Pyrene Benzonaphtofuran Benzo[a]anthracene Chrysene-D12 Chrysene PAH 252 g/mol Perylene-D12 PAH 276 g/mol

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unchanged after FL oxidation for both H and NM organic extracts. The stability of the relative intensity of oxygenated bands (especially νOH, 3,700–3,100 cm−1 and νC 0 O, 1,745–1,705 cm−1) suggests the absence of oxygenated byproducts formation. Moreover, the similarity of aliphatic profiles (νCH ali 3,000–2,800 cm −1 and δCH ali 1,470– 1,360 cm −1 ) and aromatic profiles (νCH aro , 3,100– 3,000 cm−1; νC 0 C, 1,620–1,590 cm−1; and γCHaro, 900–700 cm−1) reveals that no major molecular reorganization occurs. This lack of by-product formation was confirmed by GC-MS, where all compounds were equally degraded (Fig. 6). On the other hand, in case of AP oxidation, a small band appeared in 1,500– 1,700 cm−1 range of FTIR spectra for both H and NM extracts. Except of this oxygenated band, no significant reaction products were observed by both oxidants. Moreover, whatever the oxidation (FL or AP), nitrogen functions are preferentially removed (as in the case of H organic extract). FTIR spectra suggest the complete degradation of hydrocarbons by FL oxidation and a slight formation of oxygenated by-products or relative enrichment of initial oxygenated compounds by AP oxidation. In case of AP oxidation, GC-MS represents a relative enrichment of certain compounds (e.g., DBF, C1-DBF, and flourenone (FLO); Fig. 6, Table 1) and results in the appearance of oxygenated band in FTIR spectra (Fig. 5). Indeed, the selective behavior by AP oxidation is responsible for such relative enrichment of oxygenated compounds as observed in GC-MS chromatograms (Fig. 6; DBF, C1-DBF, FLO, and ANTDO; Table 1). On the other hand, FL oxidation resulted in equal degradation of all compounds. It is worthy to note that chemistry and mode of action of both oxidants (Fenton and persulfate) is different. In Fenton oxidation, degradation is mainly due to the attack of ·OH (Minisci et al. 1983; Legrini et al. 1993; Flotron et al. 2005; Liang et al. 2007). Hydroxyl radicals are nonselective oxidizing agents and can attack both oxygenated and nonoxygenated compounds in aliphatic and aromatic structures. Persulfate oxidation proceeds through both electron transfer and radical pathway (Minisci et al. 1983; Liang et al. 2007; Petri et al. 2011). Electron transfer reactions are relatively slow and selective as they depends upon the redox nature of compounds (oxygenated or nonoxygenated) (Petri et al. 2011).

Conclusion Ability of magnetite to catalyze Fenton reaction and activate persulfate oxidation was demonstrated in column experiments. Prior to oxidation, the availability of refractory oil residues (UCM) as well as PAHs was chemically enhanced. Significant

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degradation of the enhanced available hydrocarbons was observed, although conditions were less favorable in column compared to batch slurry system. No dissolved organic compounds were mobilized and flushed out of the system. No selective degradation was achieved by FL, while AP showed less reactivity towards high molecular weight PAHs. No by-products were observed in all oxidation systems by μ-FTIR and GC-MS. Slight decrease in reactivity was observed in column experiments as compared to batch tests. This study highlights the role of magnetite-catalyzed chemical oxidation to decrease the contamination level. Use of this approach could provide a cost-effective solution for soil remediation without pH adjustment. It can be a good alternative to the use of more than one reagent (e.g., soluble FeII + chelating agents) to catalyze oxidation at circumneutral pH. In addition, magnetite can be utilized for further oxidation cycles due to its strong structural and catalytic stability. Use of magnetite appears as a promising way to develop a remediation treatment for hydrocarbon contaminated soils. Finally, both batch and column tests suggested that the PAHs availability should be the most influencing parameter in remediation process. As PAH contaminations are often present as coal tar or asphalt and so less available for oxidation, the enhancement of PAH availability is a prerequisite to achieve an effective PAH degradation. Further work should be addressed in other kinds of soils in order to form a general conclusion. For high scale investigations (e.g., lysimeters), the transport study of magnetite and oxidant in soil columns should be also addressed.

Acknowledgments The authors gratefully acknowledge the financial support provided by the Higher Education Commission (HEC) of Pakistan and GISFI (French Scientific Interest Group–Industrial Wasteland project; www.gisfi.prd.fr).

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