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Chemical Engineering Journal 183 (2012) 357– 364

Contents lists available at SciVerse ScienceDirect

Chemical Engineering Journal

j ourna l ho mepage: www.elsev ier .com/ locate /ce j

iuron abatement in contaminated soil using Fenton-like process

ernando Vicente ∗ , Aurora Santos, Elena G. Sagüillo, Ángel M. Martínez-Villacorta, Juana María Rosas,rturo Romero

pto Ingenieria Quimica, Facultad de Ciencias Químicas, Universidad Complutense Madrid, Ciudad Universitaria S/N, 28040 Madrid, Spain

r t i c l e i n f o

rticle history:eceived 2 September 2011eceived in revised form2 December 2011ccepted 3 January 2012

eywords:itrateiuronenton-likeoil remediation

a b s t r a c t

The use of catalyzed H2O2 propagations (CHP) by the Fenton-like process has been evaluated for thetreatment of contaminated soil by diuron ((qDiuron)o = 40 mg kg−1) at natural neutral pH. The interactionof H2O2 with the soil in batch system has been previously studied at four different initial H2O2 concen-tration (20,000–40,000–60,000–120,000 mg kg−1) at room temperature in order to evaluate the oxidantdecomposition by the natural reductants species (soil organic matter and iron minerals).

The oxidation of diuron using only H2O2 at an initial concentration of 60,000 mg kg−1 was firstlydetermined (XDiuron = 55%). Another test was developed to study the effect of the addition of citrate(CCitrate = 3000 mg kg−1) as a chelating agent to increase the removal of the contaminant up to 73% sug-gesting that the concentration of iron mobilized by citrate from the soil (CFe ≈ 40 mg kg−1) promotesthe Fenton’s reaction. Finally, a third experiment was assessed using an extra amount of iron salt(CFe3+ = 600 mg kg−1) in addition to citrate, resulting to an increase in the diuron removal (XDiuron > 80%)

3+ •

because citrate makes stable iron chelates with Fe and increases the generation of OH in aqueous phaseand a higher diuron removal is obtained.

Ecotoxicity measurements in the soil and liquid phases at initial and after the Fenton’s treatment by theMicrotox® bioassay demonstrated that low toxicity is obtained in the treated soil and temporal toxicityvalues in the liquid phase is in direct correlation with the residual H2O2. The toxicity values decreasewith the H2O2 concentration to not toxic values with a final total degradation of the oxidant.

. Introduction

Herbicides are the most extensively type of pesticides used ingriculture. Due to their limited biodegradability, high vapour pres-ure and high lipid affinity, haloaromatic herbicides are considered

persistent class of chemicals [1]. Diuron is one of these herbicideshich has been frequently used in vineyards and it has been found

n natural water at concentrations exceeding the regulatory limitf 0.1 �g L−1 [2]. It is considered a priority hazardous substancey the European Commission. The dispersion of this compound ingriculture leads to pollution of the aquatic environment by soileaching [3]. Because of its toxicity for aquatic organisms and sus-icion of being carcinogenic for humans, diuron is the object of

rowing environmental concern [4].

Agricultural use of herbicides results in relatively low soil con-amination (diffuse contamination). In addition, near production

∗ Corresponding author. Tel.: +34 91 394 41 71; fax: +34 91 394 41 71.E-mail addresses: fervicen@quim.ucm.es (F. Vicente), aursan@quim.ucm.es

A. Santos), egsaguillo@quim.ucm.es (E.G. Sagüillo), angelmma@pas.ucm.es (Á.M.artínez-Villacorta), jmrosas@quim.ucm.es (J.M. Rosas), aromeros@quim.ucm.es

A. Romero).

385-8947/$ – see front matter © 2012 Elsevier B.V. All rights reserved.oi:10.1016/j.cej.2012.01.010

© 2012 Elsevier B.V. All rights reserved.

or storage of these herbicides may cause episodes of concentratedcontaminated soil caused by spillage during pesticide mixing, load-ing and rinsing operations or direct dumplign into soil disposalsites. These practices are though to be responsible for approxi-mately 45% of cases of ground water contamination [5].

In situ chemical oxidation (ISCO) seems to be a viable technologyfor remediation of pesticides. In this technique a strong chemi-cal oxidant is injected into the contaminated surface to destroythe targeted contaminants [6]. The oxidants frequently used arehydrogen peroxide (H2O2), permanganate (MnO4

−), persulphate(S2O8

2−) and ozone (O3) [6–14]. Each of these oxidants have limi-tations, (e.g., persistence, reactivity, etc.) within a soil matrix.

H2O2 has long been used in industrial applications and in watertreatment processes. It is particularly effective when it is applied incombination with a catalyst such as ferrous iron (Fe2+) to produceFenton’s reagent. The reaction chemistry is complex, but poten-tially capable of degrading a wide range of organic contaminantsdepending on conditions [15]. The following reactions may occur:

H O + Fe2+ → Fe3+ + OH− + HO• (1)

2 2

Fe3+ + H2O2 → Fe2+ + H+ + HO2• (2)

Fe3+ + HO2• → Fe2+ + O2 + H+ (3)

3 neerin

H

H

H

2

hErb

HttwipHatrdttr

tpcwtstsee2aomcoa

pdFncdto

nott

TA

58 F. Vicente et al. / Chemical Engi

O• + Fe2+ → Fe3+ + OH− (4)

O• + H2O2 → HO2• + H2O (5)

O2• + Fe2+ + H+ → Fe3+ + H2O2 (6)

H2O2 → 2H2O + O2 (7)

The combination of soluble Fe2+ salts and H2O2 producesydroxyl radicals (OH•) according to the reaction in Eq. (1). Thisquation implies a fast decomposition of H2O2 producing hydroxyladicals while the regeneration of the catalyst (Eq. (2)) is very slow,eing this last the rate controlling step.

This process of using H2O2 for ISCO has been termed “Catalyzedydrogen Peroxide Propagation” (CHP) [16]. There are also another

erms such as “Modified Fenton’s reagent” or “Fenton-like” reac-ions that have been used to refer to ISCO application with CHPhen some modifications of Classical Fenton reaction are applied

n order to solve the limitations of the classical process as acidH requierement. Modified Fenton’s reactions apply concentrated2O2 solutions [17], iron chelating agents [18–20], H2O2 stabilizinggents [21] or iron containing minerals [22–26]. H2O2 concentra-ion is higher than theoretical demand dosage required for theemediation of wastewaters by Fenton’s reagent due to the unpro-uctive consumption of the oxidant by the natural components ofhe soils. Constituents in the soil represent the bulk of interferenceshat must have to take into account to evaluate the mass of H2O2equired.

Given the various chemical structures that pesticides exhibit,he transformation pathways and mechanisms for individual com-ounds may not be well established. Thus, in cases where thehemistry associated with contaminant transformations is notell known, treatability studies will be needed. One common

heme found in studies of CHP degradation of pesticides is thatolubility and sorption degree affects to the rate and extent ofreatment achievable. In this sense higher solubility, less stronglyorbed pesticides may be degraded more rapidly and to greaterxtents [22,27]. In aqueous phase, pesticides have been morextensively studied with CHP oxidation. Oxidation of herbicides,4-dichlorophenoxyacetic acid and 2,4,5-trichlorophenoxyaceticcid in aqueous system [28–30] by CHP systems (acidic pH, lowxidant concentration, Fe2+ catalyst) often involves the release ofono-, di-, and tri-chlorophenols, after which the reaction pro-

eeds as described for chlorophenols. Little information is availablen CHP degradation of other pesticides in contaminated soil, suchs diuron.

Taking into account the above, the main objective of the studyresented here is to investigate the use of Fenton-like reaction foriuron degradation at room temperature in a sandy clay loam soil.irstly, a series of experiments were conducted with uncontami-ated soil to study the system soil-oxidant and to select the properonditions to react with contaminant of concern. Experiments wereesigned to determine the influence of various initial H2O2 concen-rations, the potential use of citrate as chelating agent and the effectf additional amount of Fe3+ as catalytic specie.

Therefore subsequent experiments were developed in contami-

ated soil to determine the effectiveness of CHP for the destructionf diuron as contaminant model by the Fenton-like process. Dueo the slow Fe3+ reduction reactions relative to the rapid Fen-on reaction (Fe2+ oxidation), a lower efficient and slower rate of

able 1verage value of the properties of the soil sample (standard deviation = 4–7%).

pH SOM (%) TC (%) TOC (%) SOM fractions IC (%)

LPI (%) LPII (%) R (%)

7.22 0.365 0.198 0.196 0.114 0.012 0.07 0.002

g Journal 183 (2012) 357– 364

HO• production occurs after Fe2+ is initially reacted with H2O2.Consequently, one disadvantage of Fe2+ amendment is that stoi-chiometric quantities are required. So Fe3+ has been selected ascatalytic specie to promote the Fenton’s reaction. Furthermore, theuse of citrate as chelating agent is required to enhance effectivenessof Fenton’s reaction under natural neutral pH conditions withoutacidifying the system. The dosage of H2O2 and citrate has beenselected according previous studies performed by the researchgroup with soils at natural neutral pH with buffer capacity [45].

2. Materials and methods

2.1. Reagents

Analytical-reagent grade were used in the experiments.Diuron 98% and sodium citrate 2-hydrate were purchased fromSigma–Aldrich. H2O2 30% (w/w) was provided from Riedel deHaën. Ferric chloride hexahydrate obtained from Fluka was usedas the catalytic Fe3+ specie. Potassium permanganate and sul-phuric acid 95–98% from Panreac were used in the determination ofH2O2. Methanol used as solvent and acetonitrile for mobile phasein HPLC were obtained from Scharlab. Naphthalene >99% fromSigma–Aldrich was used as internal standard in HPLC measure-ments. All of the suspensions and solutions were prepared withMilli-Q water (>18 M� cm) purified with a deionizing system.

2.2. Characterization of soil sample

The soil selected for this study was categorized as sandy clayloam at natural neutral pH. The properties of the soil sample arelisted in Table 1. The measurements of the properties were per-formed by triplicate to obtain a average value. Standard deviationwas between 4 and 7%. The soil selected corresponded to subsurfacehorizon (Bt) at an agriculture zone characterized by low organiccarbon content.

The pH was measured with a pH electrode (Basic 20-CRISON) in1:2.5 soil/water suspensions. The soil organic matter (SOM) contentin the soil was determined by incineration of a known weight sam-ple placed in a ceramic crucible in an electric muffle for 3 h at 550 ◦Cobtaining the corresponding SOM content by mass difference (NEN5754 Method) [31].

Soil contains two different forms of carbon: Total Organic Car-bon (TOC) and Inorganic Carbon (IC). Total carbon (TC) of the soilwas determined using a TOC analyzer as described in the standardprocedure EN 13137. Firstly, the sample was divided into two sub-samples and one sub-sample was oxidized with an oxygen stream(500 mL min−1) using TOC analyzer (TOC-V CSH Shimadzu) witha solid state module (SSM-5000A) at 900 ◦C where the formedCO2 was analyzed by non-dispersive infrared absorbance (NDIR).The second sub-sample was pre-treated with 4 mL of concentratedH3PO4 to remove the inorganic carbon (IC) and the measurementswas taken at 200 ◦C. This carbon dioxide is detected by the NDIRand the sample IC concentration is measured in the same way as

TC. The IC is a combination of carbonate and bicarbonate. The TOCconcentration is determined by difference:

TOC = TC − IC (8)

Mn (mg g−1) Fe (mg g−1) FeAmorphous

(mg g−1)FeCrystalline

(mg g−1)SBET

g (m2 g−1)

0.17 18.2 0.525 6.710 23

F. Vicente et al. / Chemical Engineering Journal 183 (2012) 357– 364 359

Table 2Operating conditions for runs carried out in batch tests. T = 20 ◦C; VL/W = 2 mL g−1.

Run Contaminant Load contaminant Initial concentration Reactant dosage

qDiuron (mg kg−1) CH2O2 (mg L−1) CFe3+ (mg L−1) CCitrate (mg L−1) CH2O2 (mg kg−1) CFe3+ (mg kg−1) CCitrate (mg kg−1)

1 Without contaminant – 10,000 – – 20,000 – –2 – 20,000 – – 40,000 – –3 – 30,000 – – 60,000 – –4 – 60,000 – – 120,000 – –5 – 30,000 – 1500 60,000 – 30006 – 30,000 300 1500 60,000 600 3000

a2wTawuwattw1bs

tCb

aalit

2

sfowaraal5wtw4

tcakamw

7 Diuron 40 30,000 –

8 40 30,000 –

9 40 30,000 300

To quantify labile and recalcitrant fractions of SOM a two stepcid hydrolysis with H2SO4 was performed. In this technique0 mL of 5 N H2SO4 was added to 0.5 g soil, and the samplesere hydrolyzed for 30 min at 105 ◦C in sealed Pyrex tubes.

he samples were centrifuged and the supernatant was takens the labile pool I (LPI). The remaining residue was hydrolyzedith 2 mL of 26 N H2SO4 at room temperature under contin-ous shaking. The concentration of the acid was then dilutedith de-ionized water to 2 N and the sample was kept for 3 h

t 105 ◦C with occasional shaking. The supernatant obtained wasaken as labile pool II (LPII) [32]. Both samples (LPI and LPII) andhe residual soil were analyzed following Walkey-Black Methodhich consist on the wet oxidation at 125 ◦C with a mixture of

N K2Cr2O7 and 96% H2SO4 [33]. The oxidation step was followedy titration of the excess dichromate solution with 0.5 N ferrousulphate.

Total Fe and Mn content in soils were determined by acid extrac-ion/atomic absorption spectroscopy (EPA 3050B Method) [34].rystalline and amorphous iron oxyhydroxides were determinedy citrate–bicarbonate–dithionite extraction [35].

The porous structure of the samples was characterized by N2dsorption–desorption at −196 ◦C, performed in a SA 3100 surfacerea analyzer (Coulter). Samples were previously outgassed for ateast 8 h at room temperature to avoid soil degradation. From the N2sotherm, the apparent surface area (SBET

g ) was determined applyinghe BET equation.

.3. Stability of H2O2 in the presence of soil

H2O2 degradation in contact with uncontaminated soil wastudied by duplicate through batch experiments. Runs were per-ormed in 20 mL glass vials, kept in continuous agitation (50 rpm)n a shaking water bath, supplied by Selecta. The temperatureas controlled and remained at 20 ◦C and the soil pH was not

djusted. The liquid phase volume to uncontaminated soil weightatio (VL/W) was 2 mL g−1. Firstly, the experiments were carried outt different initial concentrations of H2O2 (20,000–40,000–60,000nd 120,000 mg kg−1), runs 1–4 in Table 2. The samples were col-ected at different reaction times and immediately centrifuged for

min at 5000 rpm in a Centrolit Selecta centrifuge. The supernatantas analyzed for H2O2, as described in the Analytical methods sec-

ion. Each reaction vial represents one time point and all reactionsere conducted in duplicate. Vials were sacrificed at 1, 8, 24 and

8 h.Two additional experiments at the same experimental condi-

ions were performed mixing H2O2 with citrate and Fe3+/citrateomplex, runs 5 and 6 respectively showed in Table 2, to evalu-te the effect of metal chelating agent and soluble iron in the H2O2

inetic decomposition. The molar ratio of H2O2 to citrate was 176:1,nd the citrate to iron ratio was 1:1 according to the excess require-ent for Fenton’s treatment in soil system and studied in a previousork [45].

– 60,000 – –1500 60,000 – 30001500 60,000 600 3000

2.4. Fenton-like oxidation experiments

Some experiments, runs 7, 8, and 9 in Table 2, were performedby duplicate to evaluate the effectiveness of Fenton-like oxidationfor remediating contaminated soil by diuron at the same previousconditions.

Contaminated soil was artificially prepared using uncontam-inated control soil. The soil was spiked with a solution ofdiuron dissolved in methanol, initial concentration of diuronwas 100 mg L−1. Then, the mixture was completely stirred andblended to ensure homogeneity at 20 ◦C for about 72 h with aVL/W = 3 mL g−1. Finally, the mixture was laid until the methanolwas completely evaporated. During the ventilation, the mixturewas stirred each day. After the soil sample was washed withMilli-Q water with a VL/W = 3 mL g−1 stirred around 1 h to elimi-nate the residual methanol remaining in the soil. The sample wascentrifuged and the liquid phase was injected in a HPLC to deter-mine the possible diuron extracted to evaluate the loss with thewater washing. The total remaining amount of the contaminantsorbed in the soil was determined by ultrasonic solvent extrac-tion using methanol as the solvent by 3 consecutive extractions.Each step accurately weighed spiked soil was sonicated 15 min withmethanol in an ultrasonic bath (PowerSonic 505) at room tempera-ture followed by a 5 min centrifugation. The relation soil/methanolwas 1/5 g mL−1. The amount of total extracted diuron was deter-mined by HPLC. The target concentration of diuron in the spikedsoil was 40 mg kg−1.

Diuron oxidation experiments were performed in batch reactorsystems evaluating the applicability of Fenton-like. Firstly, it wasdetermined the oxidation of diuron using only H2O2 at an initialconcentration of 60,000 mg kg−1. Then, another test was devel-oped to study the effect of adding citrate (initial concentration of3000 mg kg−1) as chelating agent to extract the natural iron presentin the soil to catalyze Fenton’s reaction and to promote the removalof the contaminant. Finally, a third experiment was assessed usingan additional amount of iron salt ((CFe3+ )o = 600 mg kg−1) in addi-tion to citrate. Soil samples (1 g) were put in contact with 2 mL of thechelating agent or complex iron/citrate solution about 24 h prior tothe addition of H2O2. Then, 198 �L of the stock 30% H2O2 solutionwas added, giving a final H2O2 concentration of 60,000 mg kg−1.The Fenton-like reaction was just started upon addition of H2O2.

Vials were sacrificed at different times. Each reaction vial rep-resents one time point. In order to avoid possible explosions due togas accumulation in the vials, the caps were not sealed during reac-tion time to permit evacuation of the generated gas. The soil andaqueous phase were separated before the extraction. After centrifu-gation, the supernatant and soil extracts by ultrasonic extractionmethod previously explained were analyzed to determine resid-

ual diuron concentration by means of HPLC. Determination of theremaining H2O2 in supernatant was performed as explained below.The citrate and soluble iron in the liquid phase were also monitoredthrough the experiments. Ecotoxicity of the aqueous phase and soil

3 neering Journal 183 (2012) 357– 364

wp

2

i2csmup1Aic

tcl

t5(

cS

TbtTaiocslw

T

cr

T

r

AwctsuSR

3

An

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)o (mg·kg

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1 20 ,000

2 40 ,000

3 60 ,000

4 120 ,000

(C/C

o) H

2O

2

time (h)

60 F. Vicente et al. / Chemical Engi

ere measured by using the Microtox® bioassay to determine theotential toxic effects of the oxidation process.

.5. Analytical methods

In order to determine the residual diuron concentration andndicate the presence of its degradation products with time,0 �L of sample was injected into a high-performance liquidhromatography column. The initial concentration of the sub-trate and its transformation was monitored by HPLC (Agilent,od. 1100). A Poroshell 120 SB-C18 2.7 �m Performance col-

mn (100–4.6 mm) was used as the stationary phase. The mobilehase was acetonitrile–water (40:60, v/v) with a flow-rate off.00 mL min−1 for detection and quantification of diuron with agilent 1290 Infinity Diode Array Detector. The effluent was mon-

tored at 210 nm. Evaluation and quantification were made on ahromatography data system method (detection limit 20 ng mL−1).

The concentration of H2O2 in solution is determined by titra-ion using a solution of potassium permanganate, KMnO4, of knownoncentration, in sulphuric acid using a titrimetric volumetric ana-yzer supplied by Metrohm.

Ion chromatography was used to determine citrate concentra-ion using an anion chromatography column (Metrosep A Supp-250/40. Metrohm) and a mobile phase containing an bufferHCO3

−/CO32−) and acid suppressor.

Soluble iron concentrations were analyzed using a DR/890olorimeter with the FerroVer Iron Reagent, supplied by HACH.olution pH was measured with a pH electrode (Basic 20-CRISON).

Ecotoxicity was determined by using the Microtox® bioassay.he Microtox bioassay is based on the decrease of light emissiony Photobacterium phosphoreum resulting from its exposure to aoxicant, using a Microtox® M500 Analyzer (Azur Environmental).he inhibition of the light emitted by the bacteria was measuredfter 15 min contact time. The EC50 is defined as the effective nom-nal concentration of toxicant (mg L−1) that reduces the intensityf light emission by 50%. The parameter IC50 is defined as the per-entage of the initial volume of the sample to the volume of theample yielding, after the required dilution, a 50% reduction of theight emitted by the micro-organisms. The toxicity units (TU) of the

astewater are calculated as:

U = 100IC50

(9)

If the aqueous sample only contains a compound i at a con-entration Ci the toxicity units can be predicted by the nextelationship.

Ui = Ci

EC50i(10)

Before measuring the toxicity, the pH values of the samples weree-adjusted to between 6 and 7, in order to prevent the pH effect.

A Basic Solid Phase-Test (BSPT) was performance according thezur Environmental procedure. A 20% (w/v) solution of the soilas stirred and after 10 min while stirring its sample from the

orrect region of the beaker 2 mL of the solution and 9 serial dilu-ions were performed. The luminescence was measured beforeample addition and 30 min after the sample addition the EC50 val-es were obtained. All the chemicals used were purchased fromigma–Aldrich and the micro-organisms were Microtox® Acuteeagent supplied by I.O. Analytical.

. Results and discussion

The chemical properties of the soil studied are present in Table 1.s can be seen the soil is categorized as a sandy clay loam with aeutral pH value and low organic matter content. The labile pool

Fig. 1. Oxidant decomposition at different initial H2O2 concentrations. T = 20 ◦C;VL/W = 2 mL g−1.

fraction in organic matter represents the largest proportion of thisparameter being the easier degradable fraction (carbohydrates andproteins) [36]. The high value of crystalline iron measured impliesthat Fenton-like reagent can be successfully applied because it canact as heterogeneous catalyst, which is previously demonstrated inliterature [22–24,37].

The high specific surface value may indicate that, despite the loworganic matter content, the soil is able to retain contaminants dueto its porous structure. Furthermore, the surface area of iron-oxideminerals is also an important factor influencing the degradation oforganic pollutants by the Fenton-like reaction [38].

3.1. H2O2 decomposition by soil

H2O2 decomposition, without pH adjustment, was measured inorder to characterize the reactivity of the soil. H2O2 concentrationwas monitored with time at four different values of initial oxidantconcentration (runs 1–4 in Table 2). As can be seen in Fig. 1, H2O2decomposition depends on the initial concentration of the oxidant.When using a low H2O2 concentration a higher oxidant conversionis reached. H2O2 decomposition rates were rapid in the first hoursand decreased with time. Experiments performed on the aqueousphase without soil at 30,000 mg L−1 of H2O2 and 20 ◦C showed thatthe H2O2 concentration remains constant at 24 h, so the decreaseof the concentration is due to the interaction oxidant-soil. The nat-ural reductant species of soil such as total organic carbon (TOC) andmetallic substances can act as a limiting reagent in H2O2 decom-position and could be responsible for the “non-productive” H2O2decomposition pathways, such as the disproportion to O2 and H2O.This could explain the higher slope obtained at initial stages in Fig. 1.Otherwise the presence of iron minerals (e.g., goethite, magnetite,and hematite) or other transition metals, such as Mn, in a peroxidesystem can catalyze the hydrogen peroxide decomposition and alsofacilitate contaminant oxidation via catalysis [39–42]. The catalyticdecomposition of hydrogen peroxide due these transition metalscould explain the slope noticed at times higher than 10 h in Fig. 1.

Finally, the influence of adding a chelating agent (sodium cit-rate) and the complex Fe3+/citrate, were studied at initial H2O2concentration of 60,000 mg kg−1. The results, shown in Fig. 2, indi-cate that the influence of the citrate is somewhat comparable withthat obtained without adding chelating agent. When adding cit-rate the decomposition of H2O2 is slightly higher than the casewhen only H2O2 is added. This could be explained by the solubi-lization of iron cation from the soil to the aqueous phase due to the

chelating agent. Besides, this effect is more notorious when a salt ofFe3+ is added with citrate. The iron cation is maintained in solutionby the chelating agent at neutral pH and the catalytic decompo-sition of H2O2 increases. The reaction of Fe3+ with H2O2 leads to

F. Vicente et al. / Chemical Engineering Journal 183 (2012) 357– 364 361

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Run ( C ) (mg·kg ) (C ) (mg·kg ) (C ) (mg·kg )

3 60 ,00 0 - -

5 60 ,00 0 - 3,00 0

6 60 ,00 0 60 0 3,00 0

(C/C

o) H

2O

2

F

rr

aoa

sattem

3

T(ltTTtci

se

C

F

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7 60,00 0 - -

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9 60,00 0 60 0 3,00 0

X D

IUR

ON

(a)

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F

time (h)

ig. 2. H2O2 decomposition in soil SA for runs 3, 5 and 6. T = 20 ◦C; VL/W = 2 mL g−1.

egeneration of Fe2+ ions and subsequent formation of hydroxyladicals (OH•).

In any case, it was noticed that this soil, with low carbonatend SOM content, produces a much more slower hydrogen per-xide decomposition that the soil with higher SOM and carbonatemounts used elsewhere [43].

In order to evaluate the effects of using H2O2 as an oxidant on theoil quality, the final SOM content was measured 48 h after H2O2ddition in all the runs carried out. The results showed that ≥90% ofhe SOM initially present in the soil was degraded. It is close relatedo the highest labile (LPI) content of this soil which is constituted byasily biodegradable compounds (of which carbohydrates are theost typical).

.2. Oxidation of diuron by Fenton-like process

Operating conditions of diuron degradation are collected inable 2. The initial concentration of diuron as herbicide in soilqDiuron = 40 mg kg−1) is higher than the standard regulatory estab-ished by the RD9/2005 [44]. Initial H2O2 concentration selected forhe Fenton-like experiments to oxidize diuron was 60,000 mg kg−1.his value is expected to be enough to start the degradation process.his value is much higher than the theoretical H2O2 demand for theotal mineralization of diuron because there are non-specific lossesaused by its decomposition in the presence of reactive materialsn soils.

The theoretical H2O2 demand for the total mineralization of theorbed diuron is 109.82 mg kg−1 according to the stoichiometricquation (Eq. (11))

9H10Cl2N2O + 19H2O2 → 9CO2 + 21H2O + 2Cl− + 2NH3 (11)

Residual diuron concentration in aqueous phase and soil duringenton-like oxidation treatment at natural soil pH for soil with and

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10

20

30

40

q (

mg.k

g-1

)

time (h)

(a)

20100

Aqueo

Soil

tim

ig. 3. Residual diuron concentration in soil and aqueous phase for runs 7 (a), 8 (b) and 9

Fig. 4. Diuron conversion (a) and H2O2 decomposition (b) for runs 7–9.(qDiuron)o = 40 mg kg−1; T = 20 ◦C; VL/W = 2 mL g−1.

without chelating agent and extrinsic Fe3+ are shown in Fig. 3 (runs7–9 in Table 2). The diuron concentration decreases with time andthe removal rate of diuron is characterized by a very fast initial ratefollowed by a much slower degradation rate. As can be seen, theFenton’s reaction takes place in both phases because this may beassociated to the fast oxidation of the most accessible diuron andlower entrapped contaminant desorption to the liquid phase fromsoil. For run 7, without citrate, the residual diuron in the mediawas found in the aqueous phase as can be seen in the Fig. 3(a), andit remained constant after the first minutes. As no iron cation ispresent in this phase, the diuron desorbed is not further oxidized. Inaddition, the use of citrate (runs 8 and 9 in Fig. 3(b) and (c), respec-tively) 24 h prior to addition of H2O2 favors the diuron extractionfrom the soil to the aqueous phase improving the efficiency of Fen-ton’s reactions by a coupled desorption–oxidation mechanism [45].The extraction percentage of diuron in the liquid phase after theaddition of chelating agent and previous to the addition of the H2O2is around 25% for both runs and this factor enhances the Fenton’sreaction.

The evolution of diuron conversion in the three cases studied(runs 7–9) is compared in Fig. 4(a). Up to 55% diuron removal effi-ciency was attained in soil at 48 h reaction time with the addition

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e (h)

(b)

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(c). (qDiuron)o = 40 mg kg−1; (CH2O2 )o = 60, 000 mg kg−1; T = 20 ◦C; VL/W = 2 mL g−1.

362 F. Vicente et al. / Chemical Engineering Journal 183 (2012) 357– 364

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ig. 5. Concentration profile of citrate and Fe in aqueous phase in run 8. (CH2O2 )o =0, 000 mg kg−1; (CCitrate)o = 3,000 mg kg−1; T = 20 ◦C; VL/W = 2 mL g−1.

f H2O2. To explain the chemical oxidation in Fenton-like systemsn which Fe3+ exists in the solid phase (neutral pH), it was proposedhat H2O2 is catalyzed on the surface of iron oxyhydroxide mineralshrough surface interaction processes [39,46,47].

Efficiency in diuron removal is improved with the citrate(Ccitrate)o = 3000 mg kg−1) increasing the diuron conversion up to3%. This is because the chelating agent is capable of extractingatural ferric iron present in the soil and enhances the solubil-

ty, promoting the Fenton reaction of the diuron desorbed fromhe soil to the aqueous phase. Nevertheless the amount of Fe inhe aqueous phase is only 40 mg kg−1. Fe3+ can also be solubilizedy forming complexes with organic acid intermediates produceduring pollutant degradation [48].

As can be seen in Fig. 4(a), the higher decomposition of diurons reached in run 9 when a Fe3+/citrate complex is added. The con-ersion reached in the Fenton-like reaction is higher because moreatalyst (Fe3+/Fe2+) is available in the system. The presence of Fe3+

n the media increases the generation of hydroxyl radicals (OH•).ccording to Eqs. (1) and (2) the use of Fe3+ instead of Fe2+ pro-uces a H2O2 decomposition more controlled because the H2O2eacts rapidly with the Fe2+ and could be less effective the applica-ion of H2O2. Therefore, if results from runs 8 and 9 are compared,t can be concluded that H2O2 degradation kinetics is dependent onhe Fe cation concentration in solution.

In Fig. 4(b), decomposition of the oxidant is shown for the threeests evaluated. By comparison of Figs. 2 and 4(b), it was confirmhat the presence of pollutant (diuron in this case) in soil do notffect significantly the degradation of H2O2.

.3. Citrate and iron concentration on the reaction time

The concentrations of the chelating agent and the total Fe in theiquid phase were measured along the reaction time for runs 8 and, and the profiles are shown in the Figs. 5 and 6, respectively. As it

s shown in Fig. 5 for run 8 tested with citrate, the initial total ironxtracted from the soil by the citrate after 24 h was 39 mg kg−1,his value increases to 60 mg kg−1 after 48 h of treatment. Citrate

easured values show that the concentration of the chelating agentecreases significantly in the first hour and it remained constanturing the rest of the reaction time evaluated. This could be alsoue to an initial adsorption of the chelating agent.

On the other hand, in run 9 when a citrate and Fe solutionas added, the concentration of both reagents decrease during

he reaction time as shown in Fig. 6. In this case, the initial iron−1

oncentration is 600 mg kg , which is much higher than the run

ested with only citrate/H2O2 addition. Therefore, radicals gener-tion is higher and consequently diuron oxidation in the mediasoil–aqueous system) increases as was shown in Fig. 4.

Fig. 6. Concentration profile of citrate and Fe in aqueous phase in run9. (CH2O2 )o = 60, 000 mg kg−1; (CFe3+ )o = 600 mg kg−1; (CCitrate)o = 3000 mg kg−1;T = 20 ◦C; VL/W = 2 mL g−1.

Besides, it was noticed that citrate was continuously decreasingwith time almost disappearing completely at 50 h. Previous exper-iment in absence of H2O2 was developed in parallel showing thatno degradation in citrate and iron was noticed. This indicates thatthe removal of citrate is due to hydroxyl radicals (OH•) formed.The direct action of these radicals toward the citrate produces thedecreasing of its concentration by oxidation in the time evaluated.

This implies that the chelating agent disappears from the reac-tion media and consequently is not longer available to keep iron inthe aqueous phase at near neutral pH values to act as a catalyst.

3.4. Potential toxicity of oxidation process

Soil toxicity was evaluated in uncontaminated soil and in soilspiked with 40 mg kg−1 of diuron. The values of EC50 obtained inMicrotox assay applied to both soils were 2.87 and 1.12%, respec-tively. From these data, it is interesting to take into account thelimits proposed by Kwan and Dutka [49] for sediment toxicityassessment. Samples which produced EC50 values ≤0.5% are classi-fied as Very Toxic, 0.5 < EC50 ≤ 1% as Moderate Toxic, and EC50 > 1%as Non-Toxic. According to the values found in literature these EC50results do not imply a high toxicity in soil but they are useful tocompare the effect of diuron contamination and the impact of theremediation treatment with Fenton-like [50]. As can be seen in theresults above, the presence of the herbicide in the soil producesan increment on the soil toxicity as shows a smaller value of EC50[51,52].

At the end of the oxidation process (runs 7–9 at 48 h) soil toxic-ity including the remaining aqueous phase in pores was increasedachieving a values of EC50 = 2000 mg L−1. After several days, thistoxicity disappeared obtaining a soil toxicity similar to the uncon-taminated soil. These results are in concordance with the resultsobtained in the aqueous phase which are following described.

In Fig. 7, TU’s values in the aqueous phase were determined at48 h of the treatment for runs 7–9. The results showed that theaqueous sample with higher toxicity is that corresponding to run7. As can be seen, this simple was that containing the higher residualamount of H2O2. Lower toxicity obtained in the aqueous phase forruns 8 (addition of H2O2/citrate) or 9 (addition of H2O2/Fe3+/citrate)can be related linearly with the residual amount of H2O2 in thesamples. Diuron and H2O2 toxicity were evaluated showing an EC50value of 14 mg L−1 and 11–18 mg L−1, respectively. According to theH2O2 conversion results, the different toxicity values of these sam-ples are due to the different remaining concentration of H2O2 at48 h.

Therefore, there is a direct correlation with the H2O2 con-centration and temporal toxicity values due to amount ofherbicide/oxidation reagent ratio. It is important to point outthat this toxicity is mainly transitory, due to H2O2 decomposes

F. Vicente et al. / Chemical Engineerin

Run 9Run 8Run 70

700

1400

2100

Run (C ) (m g·kg ) (C ) (m g·kg ) (C ) (m g·kg ) *( C ) (m g·kg )

7 60,0 00 - - 41, 400

8 60 ,000 - 3,00 0 34,80 0

9 60 ,000 600 3,00 0 22,20 0

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ig. 7. TU’s values in the liquid phase at 48 h after Fenton-like treatment.qDiuron)o = 40 mg kg−1; T = 20 ◦C; VL/W = 2 mL g−1. * (CH2O2 )F : Residual H2O2 con-entration in the aqueous phase at 48 h for runs 7–9.

ompletely at 7–10 days, obtaining at the final of the treatmentery low toxicity values. The residual aqueous phase in the soilores was responsible of the remaining toxicity above cited in theoil at the end of the oxidation run (48 h).

. Conclusions

The results obtained in this study are helpful to evaluate thepplicability of ISCO using Fenton’s like for the treatment of con-aminated soils with pesticides such as diuron. The application ofHP system made possible to evaluate the effectiveness of H2O2 asxidant in the remediation of contaminated soil at natural neutralH. These results are a starting point to describe diuron oxidationith Fenton-like reagent and to analyze the toxicity associated with

he treatment. The use of citrate as chelating agent and Fe3+ asxtrinsic catalytic specie improved the results of diuron removal.

The interaction between H2O2 and soil was firstly investigatedith batch experiments to capture the overall H2O2 behavior in theresence of the soil sample. The results indicated that the decom-osition of H2O2 showed two regions, first probably related to theresence of reductant species and second with a catalytic effect dueo transition metals (i.e., Fe and Mn). The soil was characterized byow SOM but relatively high BET surface.

In the oxidation tests it has been demonstrated that the additionf citrate as chelating agent to the soil improves the efficiency of therocess into two ways: natural iron present in the soil is extractednd kept in the aqueous phase in order to act as a catalyst and diurons also desorbed being easier to degrade as it is in the aqueous phaset the same time as reactive species are generated in CHP system.

The addition of a mixture of citrate and Fe3+ to the reactionedia removed up to 80% of diuron present in aqueous–soil media.

hese are the best results achieved because the metallic specieresent in the aqueous phase is higher than in the other casesnd the homogeneous oxidation of desorbed diuron takes place atigher rate. Citrate enhances not only the iron solubilization con-ent of the soil to the liquid phase but also improves the extractionf the sorbed diuron. The pH of slurry systems is in the neutralange, reflecting the natural pH of the media. Because of the near-eutral pH, use of an Fe-chelate complex has been attempted toaintain iron availability increasing diuron removal.

The effect of the Fenton-like reaction on the microbial activity in

he soil was assessed based on toxicity tests towards Vibrio fischeri.imilar values of toxicity were obtained in the uncontaminatednd contaminated with diuron solid phase and during the Fenton

[

[

g Journal 183 (2012) 357– 364 363

oxidation process. The presence of diuron in the soil decreaseslightly the EC50 value increasing the soil toxicity. The tox-icity values achieving in the liquid phase are mainly dueto the residual oxidant present in the media and is notobserved a synergic effect with the present of the other com-ponents with no toxicity values. The toxicity values decreasewith the H2O2 concentration to not toxic values with afinal total degradation of the oxidant minimizing collateraleffects.

Acknowledgements

The authors acknowledge financial support for this researchfrom the Comunidad Autonoma de Madrid provided throughoutproject CARESOIL (S2009AMB-1648) and from Spanish Ministry ofScience and Innovation, projects CTM2006-00317 and CTM2010-16693.

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