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Chemical Engineering Journal 180 (2012) 1– 8

Contents lists available at SciVerse ScienceDirect

Chemical Engineering Journal

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

norganic chelated modified-Fenton treatment of polycyclic aromaticydrocarbon (PAH)-contaminated soils

ennya, Suyin Gana,∗, Hoon Kiat Ngb

Department of Chemical and Environmental Engineering, The University of Nottingham Malaysia Campus, Jalan Broga, 43500 Semenyih, Selangor Darul Ehsan, MalaysiaDepartment of Mechanical, Materials and Manufacturing Engineering, The University of Nottingham Malaysia Campus, Jalan Broga, 43500 Semenyih, Selangor Darul Ehsan, Malaysia

r t i c l e i n f o

rticle history:eceived 12 August 2011eceived in revised form 24 October 2011ccepted 24 October 2011

a b s t r a c t

In this work, the effects of chelating agent (CA) on the removal of phenanthrene (PHE) and fluoranthene(FLUT) from contaminated soil through modified-Fenton (MF) treatment at natural soil pH have beeninvestigated. Chelate enhanced MF treatment resulted in rapid PAH degradation that obeyed pseudo-first-order kinetics with rate constants within the ranges of 9.15 × 10−2 to 2.30 × 10−1 min−1 and 7.20 × 10−2 to

−1 −1

eywords:AHentonhelating agentydroxylersulphate

2.34 × 10 min for PHE and FLUT respectively. Inorganic CA sodium pyrophosphate (SP) was demon-strated to be superior to commonly employed organic CAs such as ethylene diamine tetraacetic acid, oxalicacid and sodium citrate, and resulted in PAH removal efficiencies of up to 89.13%. The compatibility ofSP was also tested for persulphate enhanced oxidation of PAH-contaminated soil. Remarkable oxidativePAH removals (95.41% and 92.25% for PHE and FLUT respectively) were observed in the persulphate-MFtreatment which suggested the feasibility of integrated MF reagent in treating PAH-contaminated soil.

. Introduction

PAHs are chemical compounds made up of at least two fusedromatic rings in a linear or clustered arrangement. They areydrophobic, toxic compounds characterised by high resistance toatural degradation and are commonly encountered at manufac-ured gas plants and wood treatment sites [1,2]. Soils and sedimentsften act as long term environmental sinks which tend to trap PAHolecules within the clay-humus complex [3]. On top of that, the

ccumulation of PAHs in soil matrices can lead to the contaminationf food chain which subsequently causes direct and indirect expo-ure to humans [4–8]. It has been reported that the binding forcesetween soils and PAHs vary from simple adsorption to covalentinding. Owing to the complex interactions within the soil matri-es, PAHs are resistant to hydrolysis and microbial degradation9]. Thus, this presents a major challenge in developing removalechniques for PAHs in soils.

Over the last two decades, Fenton oxidation using hydrogeneroxide (H2O2) and iron catalyst has emerged as a rapid and cost-ffective remediation solution for contaminated groundwater andoils [10]. In Fenton oxidation, the most widely accepted reaction

athway is based on the formation of hydroxyl radicals (•OH, redoxotential E0 of 2.80 V). •OH radicals are highly reactive and non-pecific oxidants [11,12]. The principal reactions occurring during

∗ Corresponding author. Tel.: +60 3 89248162; fax: +60 3 89248017.E-mail address: suyin.gan@nottingham.edu.my (S. Gan).

385-8947/$ – see front matter © 2011 Elsevier B.V. All rights reserved.oi:10.1016/j.cej.2011.10.082

© 2011 Elsevier B.V. All rights reserved.

Fenton driven transformation are listed in Eqs. (1)–(18) where RHdenotes the organic contaminant such as PAHs.

H2O2 + Fe2+ → Fe3+ + •OH + OH− k1 = 63 M−1 s−1 (1)

Fe3+ + H2O2 → Fe2+ + HO2• + H+ k2 = 2 × 10−3 M−1 s−1 (2)

•OH + Fe2+ → Fe3+ + OH− k3 = 3.2 × 108 M−1 s−1 (3)

Fe2+ + HO2• → Fe3+ + HO2

− k4 = 1.2 × 106 M−1 s−1 (4)

Fe(III) + HO2• → Fe(II) + O2 + H+ k5 = 2 × 103 M−1 s−1 (5)

•OH + H2O2 → H2O + HO2• k6 = 3.3 × 107 M−1 s−1 (6)

HO2• + H2O2 → H2O + •OH + O2 k7 = 3.1 M−1 s−1 (7)

HO2• + HO2

• → H2O + O2 k8 = 8.3 × 105 M−1 s−1 (8)

•OH + •OH → H2O2 k9 = 6 × 109 M−1 s−1 (9)

HO2• → H+ + O2

•− k10 = 1.58 × 105 M−1 s−1 (10)

H+ + O2•− → HO2

• k11 = 1 × 1010 M−1 s−1 (11)

O2•− + Fe3+ → Fe2+ + O2 k12 = 5 × 107 M−1 s−1 (12)

O2•− + Fe2+ + 2H+ → Fe3+ + H2O2 k13 = 1 × 107 M−1 s−1 (13)

HO2• + O2

•− + H2O → H2O2 + O2 + OH−

k14 = 9.7 × 107 M−1 s−1 (14)

•OH + HO2• → H2O + O2 k15 = 7.1 × 109 M−1 s−1 (15)

2 Venny et al. / Chemical Engineering Journal 180 (2012) 1– 8

Table 1Iron-chelate equilibria.

Equilibrium reaction Equilibrium constant, K

SP EDTA OA SC MA

Fe2+ + L ↔ Fe(II)L na 1014.33 Fe(II)L2:104.52 1015.5 102.5

Fe2+ + HL ↔ Fe(II)HL na 1017.2 Fe(II)L3:105.22 1019.1 naFe2+ + H+ + HL ↔ Fe(II)H2L na na na 1024.2 naFe3+ + L ↔ Fe(III)L Fe(III)H2L2:1039.2 1025.1 107.53 1025 107.1

Fe3+ + HL ↔ Fe(III)HL na 1026 Fe(III)L2:1013.64 1027.8 naH+ + L ↔ HL 108.5 1010.34 103.8 1016 104.72

H+ + HL ↔ H2L 1014.6 1016.58 105.2 1022.1 108

2H+ + HL ↔ H3L 1017.1 1019.33 na 1026.5 na1.4

N shown

•

•

R

uhiiaiItptro

twncTsrCmt

psstgoa8Nr

wtattfth

a

3H+ + HL ↔ H4L 1018.1 102

ote that for simplicity, the charges for ligand (L) and L-containing species are not

OH + O2•− → OH− + O2 k16 = 1.01 × 1010 M−1 s−1 (16)

OH + RH → R• + H2O k17 = 1 × 1010 M−1 s−1 (17)

• + H2O2 → ROH + •OH k18 = 106–108 M−1 s−1 (18)

The production of reactive •OH radicals and the ability of iron tondergo cyclic oxidation and reduction through the Fenton processave been proven to be a potentially viable approach for remediat-

ng PAH-contaminated soils [13–15]. However, the low pH (pH 2–4)nvolved in conventional Fenton oxidation often hinder practicalpplications at natural soil pH which is approximately neutral asron catalyst is easily precipitated to iron oxyhydroxides (Fe(OH)3).n order to overcome the limitation of conventional Fenton oxida-ion, chelating agents (CAs) have been introduced under natural soilH conditions to maximise catalytic activity of dissolved iron ando prevent iron lost due to binding with hydrophilic sites of natu-al organic matter [16]. This adaptation of the conventional Fentonxidation is known as the modified-Fenton (MF) treatment.

In MF treatment, complex formation reactions in theory obeyhe Lewis acid–base concept. Iron ions act as electron acceptorshich bind with chelate according to the maximum coordinationumber. CAs are characterised by their multidentate features andyclic structure that bind metal ions to form heterocyclic rings.he affinity of CA for metal ions can be defined in terms of thetability constants resulting from a series of stepwise equilibriumeactions. Table 1 summarises the stability constants (K) of selectedAs obtained from literature [17,18]. Higher log K value generallyeans that the metal ion will be bound more tightly to the CA thus

he more likely the complexes will be formed.Under near neutral pH environment, attempts to minimise iron

recipitation in Fenton systems have been made possible using CAsuch as ethylene diamine tetraacetic acid (EDTA), oxalic acid (OA),odium citrate (SC) and cathecol which are capable of enhancinghe degradation of organic contaminants such as PAHs in soils androundwater systems [15,19–21]. The performance of EDTA-MFxidation was demonstrated to be comparable to other remedi-tion approaches such as persulphate oxidation under which up to0% degradation of total PAHs in field soil could be achieved [22].evertheless, EDTA itself poses environmental risks because of its

efractory nature and persistency in the environment [23].Recently, sodium persulphate (Na2S2O8) or simply persulphate

as introduced as a potential oxidant for in situ soil and groundwa-er remediation [22,24,25]. Persulphate anion (S2O8

2−) can remainctive longer than conventional Fenton’s reagent after delivery intohe soil matrix [26]. S2O8

2− is a strong oxidant with a redox poten-ial of 2.01 V. It can be activated by transition metal (Mn+) such aserrous (Fe2+) and ferric (Fe3+) ions to initiate a radical pathway

hrough the formation of sulphate radical anion (SO4

•−) with evenigher redox potential (E0 = 2.60 V).

The generation of SO4•− can be accelerated at ambient temper-

ture as described in Eq. (19). In the presence of excess metal ions,

na 1029.5 na

(EDTA:L4− , OA:L2− , SC:L4− , MA:L2− and SP:L4−). na: data not available [26,27].

further reaction of the metal ions with SO4•− as shown in Eq. (20)

may take place.

S2O82− + Mn+ → SO4

•− + SO42− + Mn+1 (19)

SO4•− + Mn+ → SO4

2− + Mn+1 (20)

Over the last 10 years, iron activated persulphate technologyhas drawn much interest for its capability in destroying organicpollutants (trichloroethylene, pentachlorophenol, BTEX and PAHs)in soil systems. The iron activated persulphate oxidation has beenassessed in laboratory-scale experiments with single and sequen-tial addition of iron catalyst, chelated with either EDTA or SC[24,26–29]. In addition, the relatively high stability of activatedpersulphate in soil subsurface under normal conditions and theformation of benign end products have resulted in persulphate oxi-dation being a promising alternative for the enhancement of soilremediation [30,31].

Due to the fact that the works on iron chelate applications arelimited to the use of organic CAs such as EDTA, OA and SC, this workaims to investigate the feasibility of MF reagent coupled with inor-ganic CA for PAH oxidation in soil systems, an area of research whichhas yet to be studied and reported. The advantages of using inor-ganic CA include (1) likely lesser competition for generated activeradicals responsible for pollutant oxidation and (2) total organiccarbon is not increased throughout the course of the pollutantdegradation [21]. Previous findings have revealed that inorganic CApyrophosphate exhibited a positive effect in groundwater remedia-tion of trichloroethylene at neutral pH [32]. Besides, phosphate ionsmay serve as a nutrient source to the soil matrix and its microorgan-isms. Therefore, the efficiency of sodium pyrophosphate (SP) as aninorganic CA representative in remediating PAH-contaminated soilis evaluated and compared to commonly employed CAs includingEDTA, OA, SC and malic acid (MA). In line with the recent emer-gence of persulphate oxidation, the compatibility of persulphateenhanced MF oxidation chelated with SP is also assessed here toidentify the future direction of this research area.

2. Materials and methods

2.1. Chemicals

Phenanthrene (PHE, 90%) and fluoranthene (FLUT, 98%) werepurchased from Sigma–Aldrich. H2O2 (30%) and sodium persul-phate (Na2S2O8, 99%) were purchased from Merck while ferricsulphate (Fe2(SO4)3·xH2O, 97%) was from Sigma–Aldrich. EDTA, OA,MA and SP were purchased from Merck and SC was from FisherScientific.

Dichloromethane (DCM, 99.5%, AR analysis), acetone (99.5%, ARanalysis), calcium chloride dehydrate (CaCl2, 99 +%, ACS grade),sulphuric acid (H2SO4, 98%), n-hexane (≥96%) and mercury chlo-ride (HgCl2, >99.5%) were purchased from Merck. Acetonitrile (ACN,

Venny et al. / Chemical Engineering Journal 180 (2012) 1– 8 3

Table 2Chelate enhanced MF oxidation and reactant dosages.

CA Experiment 30% H2O2 (mL) 1 M Fe3+ (mL) CA (g) H2O2:soil:Fe3+:CA (w:w:w:w)

SP MFa1 0.735 0.9 0.4015 0.049:1:0.072:0.080MFa2 0.735 0.9 0.2007 0.049:1:0.072:0.040MFa3 0.735 0.9 0.1338 0.049:1:0.072:0.027

EDTA MFb1 0.735 0.9 0.3350 0.049:1:0.072:0.067MFb2 0.735 0.9 0.1675 0.049:1:0.072:0.034MFb3 0.735 0.9 0.1117 0.049:1:0.072:0.022

OA MFc1 0.735 0.9 0.1135 0.049:1:0.072:0.023MFc2 0.735 0.9 0.0567 0.049:1:0.072:0.011MFc3 0.735 0.9 0.0378 0.049:1:0.072:0.008

SC MFd1 0.735 0.9 0.2647 0.049:1:0.072:0.053MFd2 0.735 0.9 0.1323 0.049:1:0.072:0.026MFd3 0.735 0.9 0.0882 0.049:1:0.072:0.018

MA MFe1 0.735 0.9 0.1207 0.049:1:0.072:0.024MFe2 0.735 0.9 0.0603 0.049:1:0.072:0.012

0.9

5

9Rp

2

MGpgm

2

Bdu(

P

b(1t

foawTt[

nEoM(prtwa

MFe3 0.735

g of soil, 480 mM H2O2, 60 mM Fe3+, solution:soil = 15 mL:5 g.

9.8%, HPLC grade) and n-pentane (99%, R&M) were purchased fromank Synergy. Anhydrous sodium sulphate (Na2SO4, 99 +%) wasurchased from Fisher Scientific.

.2. Soil samples

Surface soil samples (0–10 cm) were collected from Semenyih,alaysia (N 02◦ 56′ 48.8′′, E 101◦ 52′ 32.7′′, 34 m above sea level).ummy and fibrous materials which were not part of the soil sam-les (e.g. leaves) were removed since they were not amenable forrinding. The soil samples were air-dried and passed through 2 mmesh using laboratory test sieve (Endecotts Ltd., England).

.3. Soil characterisation

The particle size analyses were determined according to theuoyoucos hydrometer method [33]. The soil bulk density wasetermined using an oven-dry (130 ◦C for 20 h) basis per unit vol-me [33]. Subsequently, the soil porosity was estimated using Eq.21) assuming a particle density of 2.65 g cm−3.

orosity (%) =(

1 − bulk densityparticle density

)× 100% (21)

Meanwhile, the specific surface area of soils was determinedy the BET nitrogen adsorption method with a porosimeterMicromeritics, ASAP 2020). A degassing vacuum condition of.33 Pa for 3 h and degassing temperature of 150 ◦C were appliedo determine the specific surface area of the soil sample [34].

The measurement of moisture content for the fresh soil was per-ormed using a moisture analyser (MX-50, A & D Company) whichperated at 105 ◦C and 0.05%/min. The pHH2O test was conductedccording to the US EPA Method 9045D whereas the pHCaCl2 testas conducted by replacing the amount of water with 0.01 M CaCl2.

he measurement of soil organic matter (SOM) was conducted viahe Walkley–Black method and the loss-on-ignition (LOI) method35].

The total iron available in the soil sample was measured byitric acid digestion. The experimental procedures followed the USPA Method 3050B. In brief, 2 g of dried soil was digested with-ut boiling at 95 ± 5 ◦C in a 250 mL digestion vessel (RotamantleS-ES) with repeated additions of 10 mL concentrated nitric acid

1:1, v/v, in water) and 3 mL of H2O2. The reaction was allowed toroceed until no brown fumes were given off indicating complete

eaction with the concentrated nitric acid. Subsequently, the diges-ate was filtered through Whatman No. 41 filter paper. The filtrateas then diluted with distilled water and analysed using an atomic

bsorption spectrometer (Perkin–Elmer Analyst 400).

0.0402 0.049:1:0.072:0.008

2.4. Spiking of soil samples

PAHs were spiked onto the matrix of the collected uncontami-nated soil. The spiking procedure was a modified method adaptedfrom Northcott and Jones [36]. A 30 g soil sample (dried and sievedto 2 mm mesh) was divided into 6 portions. Each portion of 5 guncontaminated soil was placed in a glass vial and 500 �L of PAHs-DCM stock solution was poured to make up a total concentration of1000 mg/kg dry soil containing PHE and FLUT. Despite the fact thatmany authors have used lower levels of PAHs in spiked soil, e.g.100 mg/kg to 300 mg/kg [37–40], elevated contaminant concentra-tion and multiple PAHs were used in this study to better reflect thePAH load in contaminated field soils.

The stock solution was shaken using a Vortex Mixer (REAX, C/N541, Heidolph) for 1 min prior to mixing with the fresh soil. Thestock solution resulting in 500 mg/kg each of PHE and FLUT wascarefully added onto the soil matrix surface to avoid PAH loses onthe recipient walls or to the base of the vials. After the stock solutionwas added to the fresh soil, mixing was performed thoroughly usinga stainless steel spatula for 1 min. The procedures were repeated forthe remaining portions and the solvent was allowed to evaporatefor 3 h in a fume hood. All portions were mixed and transferred toa clean glass vial.

The spiked soil samples were subsequently aged for 15 daysin a refrigerated environment (4 ◦C) [41], in order to prevent soilmicrobes to remain active and continue to degrade labile com-pounds.

2.5. PAH oxidation by CA enhanced MF treatment

Following contaminant aging, the soil samples were subjectedto a slurry phase treatment with Fenton’s reagent in 100 mL borosil-icate vials with screw caps. A 5 g sample of spiked soil was slurriedin 15 mL of distilled water containing a specific amount of CA(0.1135–0.4015 g) and 0.03 g of HgCl2 (0.2%, w/v) in order to inhibitpotential microbial activity [15,42] as listed in Table 2.

Rapid decomposition of the H2O2 as an oxidant in treatmentscoupled with Fe2+ ions often hinder the effective application of Fen-ton’s reagent in removing PAHs from soil matrices. For this reason,it is necessary to ensure prolonged lifetime of H2O2 in the soil slurrytreatment. Therefore, Fe3+ was employed instead of Fe2+ to avoidinstability of the oxidant H2O2 when reacted with Fe2+. The MFtreatment coupled with either inorganic or organic chelated Fe3+

was carried out at natural soil pH. The Fenton oxidation was initi-ated upon the addition of 0.735 mL of 30% H2O2 and the reactiontook place in a hooded horizontal water bath shaker (MemmertWNB 7-45, Germany) operated at 30 ± 0.1 ◦C and 150 strokes per

4 Venny et al. / Chemical Engineeri

Addition of CA solution containing0.2 % (w/v) HgCl2 into 5 g of soil

samples

Addition of 0.9 mL of 1 M Fe3+

Gradual addition of 30 % H2O2and recording of reaction time

MF oxidation in water bath shakeroperated at 30 oC

Filtration

Solid phase:Automated Soxhlet

extraction

Evaporation of extract

Analysis of PAH compounds

Soil collection

Drying and sievingto 2 mm mesh

Aqueous phase:Liquid-liquid

extraction

PAH aging of 15 days

m6tmpctom

2

nowo(NatWmw

by gas chromatography

Fig. 1. Flow chart of MF treatment.

inute. At various time intervals (5 min, 10 min, 15 min, 30 min,0 min, 120 min, 180 min and 24 h), samples were withdrawn fromhe shaker, treated with five drops of 1 N H2SO4 (pH ≈ 1) to ter-

inate the Fenton oxidation and subjected to phase separationrocess. Control experiments were also carried out under the sameonditions without the Fenton’s reagent i.e. the volume of the Fen-on’s reagent (H2O2 and Fe3+) was substituted by distilled water. Allf the MF oxidation experiments were performed using H2O2: Fe3+

olar ratio of 8:1 (n = 3). Fig. 1 is a flow chart of the MF treatment.

.6. PAH oxidation by persulphate oxidation

The oxidation processes were conducted with different combi-ation of oxidants (persulphate only, H2O2 activated persulphater persulphate enhanced MF oxidation) in 100 mL borosilicate vialsith screw caps. A 5 g sample of spiked soil was slurried in 10 mL

f distilled water containing 0.4015 g of SP and 0.03 g of HgCl20.2%, w/v). This was then followed by the addition of 0.43 g ofa2S2O8 (dissolved in 5 mL distilled water), with or without theddition of 0.5 mL H2O2 and 0.9 mL of Fe3+. The Fenton reaction

ook place in a hooded horizontal water bath shaker (Memmert

NB 7-45, Germany) operated at 30 ± 0.1 ◦C and 150 strokes perinute. At various time intervals (10 min, 30 min and 24 h), samplesere withdrawn from the shaker and subjected to phase separation

ng Journal 180 (2012) 1– 8

process. The persulphate based experiments were carried out atnatural soil pH following different combination of reactant dosagesas shown in Table 3 (n = 3).

2.7. Analytical methods

2.7.1. PAH extractionsAfter the soil slurry oxidation treatment, the supernatant was

separated from the solid phase by vacuum filtering with filter paper(Sartorius Stedim Biotech 389). For PAH extraction from the solidphase, automated Soxhlet extraction (Gerhardt Soxtherm) wasselected because it is not as time consuming as other extractionswhile providing comparable extraction efficiency [43]. The sepa-rated solid phase was placed in an extraction thimble (Advantec,ID (84) 33 mm × 80 mm, Japan) and mixed with Na2SO4 (initiallypurified by heating at 400 ◦C for 4 h) at a ratio of 2:1 (w/w) (10 g per5 g of soil sample) to reduce the moisture level. The extraction wasperformed according to the US EPA Method 3540C. The PAHs in theaqueous phase were extracted by means of liquid–liquid extractionusing n-hexane at a ratio of 1:1 (v/v).

2.7.2. Gas chromatography (GC) analysisThe PAHs from both solid and aqueous extracts were analysed

using a GC (Clarus 500 Agilent USA), equipped with a flame ion-isation detector (FID) and fused silica capillary column (DB-5MS,30 m × 0.25 mm × 0.25 �m), according to the US EPA Method 8100.Helium gas (110.32 Pa) was used as the carrier gas. The tempera-tures of injector and detector were 290 ◦C and 300 ◦C respectively.The temperature in the oven was set at 100 ◦C for 1 min, ramped at25 ◦C/min to 310 ◦C and held for 2 min. The concentration of individ-ual PAH in the solvent was quantified using the external standardcalibration method with 5-point standard curves (all R2 > 0.98). Thelimits of detection were found to be 1.24 and 0.9 mg/kg while thelimits of quantification were 1.97 and 1.30 mg/kg for PHE and FLUTrespectively. Under the specified operating conditions, individualPAHs were identified by retention times of 8.2 min and 9.6 min forPHE and FLUT respectively.

2.7.3. H2O2 decomposition analysisIn order to investigate H2O2 decomposition, a soil sample was

sacrificed at specific time intervals between 5 min and 3 h. The soilslurry was subsequently centrifuged at a speed of 7000 rpm for3 min. The concentration of residual H2O2 (1 mL supernatant) wasdetermined from reduction-oxidation titration using potassiumpermanganate (KMnO4). Stoichiometrically, at ambient temper-ature of 30 ◦C, H2O2 reacts with permanganate according to thefollowing: 2MnO4

− + 5H2O2 + 6H+ → 2Mn2+ + 5O2 + 8H2O [44].

3. Results and discussion

3.1. Soil physicochemical properties

The soil particles were mainly present within the range of0.15–0.60 mm. The soil classification and its physicochemical prop-erties are presented in Table 4. These include the measurementof two important properties, namely SOM and naturally occur-ring iron. The former often results in the scavenging effect towards•OH radicals designated for PAH oxidation while the latter plays animportant catalytic role in H2O2 decomposition.

3.2. Performance of chelate enhanced MF oxidation

3.2.1. H2O2 decomposition and PAH degradation kineticsAs shown in Fig. 2, the decomposition of H2O2 took

place very rapidly during the chelate enhanced MF oxidation

Venny et al. / Chemical Engineering Journal 180 (2012) 1– 8 5

Table 3Persulphate based oxidation and reactant dosages.

Experiment Na2S2O8/Per (g) 30% H2O2 (mL) 1 M Fe3+ (mL) SP (g) Per:H2O2:Fe3+:soil (w:w:w:w) Solution:soil (mL:g)

P 0.43 – – 0.4015 0.086:0:0:1 3:1PHP 0.43 0.5 – 0.4015 0.086:0.033:0:1 3:1MFPa 0.43 0.5 0.9 0.4015 0.086:0.033:0.072:1 3:1

5 g of soil, Fe3+:SP = 1:1 mol:mol, a Na2S2O8:Fe3+ = 2:1 mol:mol.

Table 4Soil sample characterisation and properties.

Characteristic Loamy sand

Sand (%) 85.68Silt (%) 1.44Clay (%) 12.88Bulk density (g/mL)a 1.33 ± 0.02Specific surface area (m2/g soil)b 2.23 ± 0.25Porosity (%)a 49.76 ± 0.79Moisture content (%)a 0.34 ± 0.03pH in water at 25.8 ± 0.1 ◦Ca 7.13 ± 0.02pH in 0.01 M CaCl2 at 23.8 ± 0.1 ◦Ca 6.28 ± 0.01Walkley–Black SOM (%) 0.20Loss on ignition SOM (%)a 0.26 ± 0.07Total iron (mg/g soil)b 29.84 ± 0.01

(tfcr0lhlupa

corrtP

d

Fcfm

0.00

0.50

1.00

1.50

2.00

2.50

3.00

3.50

200150100500

-ln (C

t/Co)

Reaction time (min)

MFa1-PHE

MFa1-FLUT

MFa2-PHE

MFa2-FLUT

MFa3-PHE

MFa3-FLUT

Fig. 3. PHE and FLUT degradation kinetics for SP-MF treatment at different Fe3+:SP

a Average of 4 replicate determinations.b Average of duplicate determinations.

Fe3+:CA = 1:1 mol:mol) indicating that chelated Fe3+ in the solu-ion reacted with the H2O2 yielding •OH radicals within a short timerame (H2O2 half-life varied from 5 min to 20 min), except in thease of SP. Note that the results presented took into account the PAHesidual concentration in the aqueous phase (0.27–4.14 mg/kg and.61–7.27 mg/kg for PHE and FLUT respectively). The occurrence of

ower H2O2 decomposition rate in MF1 could be attributed to (1)igher stability of the oxidant (H2O2) when mixed with SP or (2)

ess scavenging effect contributed by SP during the MF treatmentsing inorganic CA. This phenomenon in turn resulted in less com-etition for generated •OH radicals responsible for PAH oxidationnd lower oxidant demand (reduced operating cost).

It is worth noting that although Fig. 2 shows that organic CAsould give rise to rapid generation of •OH radicals at the beginningf the treatment, the scavenging effects from simultaneous sideeactions might have taken place and limited the availability of •OHadicals for PAH oxidation (Eqs. (3), (6), (9), (15) and (16)) becausehe MF reagents would possibly require longer time to oxidise the

AHs adsorbed onto the soil matrix.

In parallel to the decomposition of H2O2, both PHE and FLUTegraded rapidly within the first 10 min but after 30 min, PAH

0.00.10.20.30.40.50.60.70.80.91.0

200150100500

H2O

2re

sidu

al (C

/Co)

Reaction time (min)

SP

EDTA

OA

SC

MA

ig. 2. H2O2 residual for MF treatment in the presence of organic (SP) and inorganichelating agents (EDTA, OA, SC and MA). Error bars of triplicates are not shownor clarity. Solid lines are exponential decay trendlines which do not represent any

odel.

molar ratios. Nomenclature for experiment runs as given in Table 2.

degradation slowed down considerably and reached a plateau. Itis obvious from Figs. 3 and 4 that the degradation of PAHs using SPand EDTA followed a two stage process. Other organic CAs exhibitedsimilar PAH degradation behaviour (data not shown here) which isin agreement with the results reported in the literature for the oxi-dation of other aromatic hydrocarbons such as BTEX and phenolusing Fenton process [45,46]. The first stage occurred within thefirst 10 min while the second stage started from after 10 min tocompletion. Therefore, such degradation cannot be described by asingle rate expression for the complete course of the treatment.

The degradation kinetics of PAH compounds towards •OH rad-icals can be expressed by a second order reaction, as shown in Eq.(22) where [PAH] represents the PAH concentration at time t inmg/kg.

−d [PAH]dt

= k17[•OH]SS [PAH] (22)

Since free radical reactions are very fast, it is assumed that asteady-state •OH concentration was achieved as soon as the Fenton

reaction started. Therefore, the steady-state •OH concentration was

0.00

0.50

1.00

1.50

2.00

2.50

200150100500

-ln (C

t/Co)

Reaction time (min)

MFb1-PHE

MFb1-FLUT

MFb2-PHE

MFb2-FLUT

MFb3-PHE

MFb3-FLUT

Fig. 4. PHE and FLUT degradation kinetics for EDTA-MF treatment at differentFe3+:EDTA molar ratios. Nomenclature for experiment runs as given in Table 2.

6 Venny et al. / Chemical Engineeri

Table 5Pseudo-first-order rate constants using inorganic and organic CAs.

CA Experiment PAH k′ (min−1) R2 RSD (%)

SP MFa1 PHE 0.2295 0.87 9.81FLUT 0.2257 0.95 18.28

MFa2 PHE 0.2932 0.96 19.86FLUT 0.2355 0.90 15.64

MFa3 PHE 0.1466 0.96 15.81FLUT 0.1167 0.91 9.29

EDTA MFb1 PHE 0.1795 0.98 14.85FLUT 0.1456 0.96 7.77

MFb2 PHE 0.1696 1.00 16.29FLUT 0.1547 1.00 6.63

MFb3 PHE 0.1096 0.99 10.38FLUT 0.1064 0.98 18.35

OA MFc1 PHE 0.1004 0.86 11.39FLUT 0.0946 0.84 5.14

MFc2 PHE 0.1829 0.99 6.38FLUT 0.1379 0.92 4.90

MFc3 PHE 0.1382 0.99 9.34FLUT 0.1279 1.00 5.07

SC MFd1 PHE 0.0936 0.99 18.78FLUT 0.0821 0.91 9.94

MFd2 PHE 0.0981 0.85 1.15FLUT 0.0836 0.89 1.83

MFd3 PHE 0.0915 0.95 2.91FLUT 0.0720 1.00 9.28

MA MFe1 PHE 0.1407 0.95 10.04FLUT 0.1440 0.90 12.35

MFe2 PHE 0.1392 0.88 16.96FLUT 0.1291 0.96 3.39

MFe3 PHE 0.1376 0.95 9.60

ao

−

wsroitfi

dtavvrec[

FLUT 0.1030 0.99 6.48

ssumed to be constant and thus Eq. (22) simplifies to pseudo-first-rder kinetics as shown in Eq. (23).

d[PAH]dt

= k′ [PAH] (23)

here k ’ = k17[• OH]SS denotes the pseudo-first-order rate con-tant. For all types of CAs tested, the normalised experimental dataevealed good linearity (correlation coefficient, R2 = 0.84–1.00) thatbeyed pseudo-first-order kinetics. Owing to the fact that •OH rad-cals have a very short lifespan in aqueous solution [45,47,12] onlyhe data from the first rapid stage were used to obtain the pseudo-rst-order rate constants listed in Table 5.

The kinetics analysis revealed that Fe3+:CA ratio was notirectly proportional to the reaction rates. For instance, MFa2 hadhe highest rate constant compared to MFa1 and MFa3. Over-ll, the pseudo-first-order rate constants for PHE degradationaried within 9.15 × 10−2 to 2.30 × 10−1 min−1 while for FLUTaried within 7.20 × 10−2 to 2.34 × 10−1 min−1. Kanel et al. [39]

eported that heterogeneous oxidation of PHE in sand in the pres-nce of goethite followed pseudo-first-order kinetics with rateonstants within 2 × 10−4 to 1.1 × 10−3 min−1. Valderrama et al.48] also described that the Fenton oxidation kinetics of PAHs

61.6755.79

41.7331.09

44.58

17.86 22.89 31.37

27.4423.95

0102030405060708090

100

MASCOAEDTASP

PAH

rem

oval

(%)

24 h30 min

a b

Fig. 5. PAH removal for chelate enhanced

ng Journal 180 (2012) 1– 8

in aged soil samples with creosote oil from a wood preservingsite followed pseudo-first-order. The reported rate constants were6.5 × 10−4 min−1 and 3.33 × 10−4 min−1 for 3-aromatic ring PAH(e.g. PHE) and 4-aromatic ring PAH (e.g. FLUT) respectively.

Noticeably, the values reported in the literature are a few ordersof magnitude lower compared to the values obtained here but inthe previous studies, the iron catalysts used and other operatingparameters were different. No chelating agents were utilised andin some cases, aged or field soils were used instead of spiked soils.It is important to also take note that native PAHs are often morepersistent than spiked PAHs due to their strong sorption onto thesoil matrix [15,48,49].

3.2.2. Chelation efficacy towards PAH removalFig. 5 depicts similar trends of PHE and FLUT removals regardless

of the type of CA employed to the MF treatment of PAH-contaminated soil. By extending the reaction time from 30 minto 24 h, PHE removal increased from 61.70% to 79.53%, 55.79% to78.68%, 41.73% to 73.10%, 31.09% to 58.53% and 44.58% to 68.53%in the case of SP, EDTA, OA, SC and MA respectively. Likewise, FLUTremoval also improved by 18.62%, 24%, 31.8%, 26.51% and 24.13% inthe case of SP, EDTA, OA, SC and MA respectively. The slower rate ofPAH degradation at the end of the reaction could be attributed to (1)lower concentration of available PAHs in the aqueous phase for fur-ther oxidation, (2) limited •OH radicals produced when H2O2 wasnearly completely decomposed, (3) complexation of iron chelatewhich reached equilibrium state thus controlling the amount ofsoluble iron catalyst in the soil slurry and (4) desorption of PAHstightly associated with the soil matrix.

This hypothesis is particularly true in the case of SP, EDTA, OAand MA chelated treatments in which the logarithmic of the sta-bility constants for the CA-Fe3+ chelates formation (Table 1) arein decreasing order (39.2, 25.1, 7.53 and 7.1 respectively). Follow-ing this sequence, EDTA-Fe3+ complex was the most tightly boundamong the tested organic CAs and was second only to SP-Fe3+

chelated system. While the logarithmic of SC-Fe3+ chelate stabilityconstant is reported to be approximately 25 (between EDTA andOA), the SC-Fe3+ MF treatment did not appear to be more effec-tive than the OA- and MA-Fe3+ MF treatments. A likely reason forthis could be the steric factors of the SC-Fe3+ chelate formation,i.e. despite the necessary number of donor atoms being present,chelate formation may fail to take place at all or only to a smallextent since the ring formation is completely or partially hindered[17].

Apart from that, it was noted that a relatively higher rate of H2O2decomposition resulting from the use of EDTA, SC and MA (Fig. 2)did not necessarily reflect a higher PAH removal efficiency com-

pared to SP and OA (Fig. 5). Therefore, it is reasonable to concludethat the presence of scavenging effect from SOM and other compo-nents within the soil slurry might have inhibited the oxidation ofPAHs by •OH radicals formed during the MF treatment.

70.50

54.4540.90

30.69

51.91

18.6324.06 31.84

26.51

24.13

0102030405060708090

100

MASCOAEDTASP

PAH

rem

oval

(%)

24 h30 min

MF treatment. (a) PHE and (b) FLUT.

Venny et al. / Chemical Engineeri

81.48 83.71 92.2 0 66.00 63.22 86.73

7.70 9.37 0.41

17.61 23.85

0.480.96 0.44 2.80

0.32 0.175.04

0102030405060708090

100PA

H re

mov

al (%

)

PHE (3-aromatic ring)

24 h

30 min

10 min

FLUT (4-aromatic ring)

Fp

reSi(

ermspmh

tmttaCrrtct(oaacct

atsoFctm(

3

si

ig. 6. PAH removal for persulphate based treatment (persulphate only (P),ersulphate-H2O2 (PHP) and integration of modified Fenton and persulphate (MFP)).

Among all the CAs employed, SP resulted in the highest PAHemoval efficiency for both PHE and FLUT, with the followingfficiency sequence of SP > EDTA > OA > MA > SC. The application ofP-MF treatment on the contaminated soil has verified the capabil-ty of inorganic CA to oxidise PAHs at remarkably low concentration[SP] = 60 mM; Fe3+:SP = 1:1 mol:mol).

On the other hand, EDTA showed the highest PAH removalfficiency among the tested organic CAs for both PHE and FLUTemovals. This could be linked to the intrinsically strong five-embered ring structure in EDTA-iron complex which is highly

table even at natural soil pH environment. However, it has beenreviously reported that the persistency of EDTA in the environ-ent can contribute to heavy metal bioavailability and public

azard [21].In comparison to EDTA, the other organic CAs tested form rela-

ively weak complexes with Fe3+ ions. Both PHE and FLUT degradedore slowly but to a similar extent by both OA- and MA-MF

reatments. Despite the fact that SC is a biodegradable CA, SC-MFreatment resulted in the lowest PAH removal efficiency (58.53%nd 57.52% for PHE and FLUT respectively) compared to all testedAs. Similar observation of the slow H2O2 dissociation and PAHemoval associated with SC chelating ability has been previouslyeported for the oxidation of pentachlorophenol in groundwa-er system [50]. In addition, SC was reported as a relatively pooratalyst of the Haber-Weiss cycle [51]. This was presumably dueo high reaction rate between this organic acid and •OH radicalskSC/

•OH = 3.2 × 108 M−1 s−1) at pH 6.6 [50]. Despite the lower rate

f PAH removal observed throughout the SC-MF treatment, gradualnd continuous increase in PAH degradation when the reaction wasllowed to proceed to 24 h suggested that the application of SC-Fe3+

helates is feasible for slow activation of H2O2 in removing recal-itrant pollutants such as PAHs which needs a prolonged reactionime for PAH desorption and oxidation from the soil matrix [21].

Despite its lower efficiency in degrading PAHs, MA which is more biodegradable CA than EDTA has been successfully usedo form complexes with Fe(III) for sulphur nanoparticles synthe-is from hydrogen sulphide [52]. In the present study, it wasbserved that PHE had a greater reduction in concentration thanLUT (in general less susceptible to degradation) except for theases when SP or MA was applied as the CA. These findings advocatehe advantages of using SP and MA to effectively degrade higher

olecular weight (HMW) PAHs of considerably high concentration500 mg/kg soil).

.2.3. Persulphate based chemical oxidationThe main results of the experimental investigation of per-

ulphate based chemical oxidation are presented in Fig. 6. Asllustrated in Fig. 6, the experiment performed with persulphate

ng Journal 180 (2012) 1– 8 7

as the only oxidant (P) resulted in PAH removal efficiencies ofup to 90.14% and 83.93% for PHE and FLUT respectively after24 h oxidation. Following the addition of H2O2 into the persul-phate based oxidation (PHP), the PAH removal efficiencies slightlyimproved to 93.52% and 87.24% for PHE and FLUT respectively,indicating only little influence of H2O2 on the persulphate basedtreatment. It is likely that more active radicals were generated inthe presence of persulphate and H2O2 causing competing effectsbetween SO4

•− and •OH radicals for PAH oxidation which in turngave only minor improvement in PAH removal. In comparison,persulphate-MF treatment (MFP) destroyed up to 95.41% of PHEand 92.25% of FLUT. These findings imply the compatibility ofSO4

•− radical anion (Eqs. (19) and (20)) in enhancing the MFtreatment coupled with SP even at remarkably low concentra-tion (Na2S2O8:Fe3+ = 2:1 mol:mol) compared to reported values forremediation of BTEX and PAHs in contaminated soils [24]. Onepossible explanation is that the SO4

•− generated from the decom-position of persulphate eventually lowered the pH of the solutionin the presence of Fe3+ solution which is acidic in nature (in thiscase pH ≈ 2.6–3). Unlike the experiments P and PHP, the PAH resid-ual concentrations from MFP could conform to the targeted limitof 40–50 mg/kg [45], i.e. 22.95 mg/kg and 38.75 mg/kg for PHE andFLUT respectively.

It is noteworthy that persulphate based oxidation was foundto be more effective in treating PHE from the soil samples. Theaddition of persulphate is therefore more recommended for theremediation of soils contaminated with low molecular weightPAHs. Overall, the excellent performance of persulphate enhancedMF reagent for the soil sample of concerned suggests the fea-sibility of integrated MF reagent (MFP) for the remediation ofPAH-contaminated soils.

3.3. Associated treatment costs and soil quality criteria

For all the tested CAs, the H2O2 requirement which often deter-mines the major cost of Fenton treatment was low, i.e. 480 mMH2O2 per 5 g of soil samples. This was accompanied by very lowconcentration of Fe3+ catalyst (i.e. 0.9 mmol per 5 g of soil sam-ple), relatively low concentration of CA (i.e. Fe3+:CA = 1:1 mol:mol)and a mild concentration of persulphate for the MFP experi-ments resulting in Na2S2O8:Fe3+ = 2:1 mol:mol compared to similarworks reported in the literature [28,30,29]. Furthermore, the lowmain reactant dosages applied also suggested appreciable costsreduction relative to other soil slurry MF experiments [22] (i.e.100–200 mmol of H2O2 per 30 g of sample with resulting concen-tration of 1–2 M H2O2).

Apart from that, SP as the proposed inorganic CA could fulfilthe soil quality criteria and conformed to the Ontario, British andCanadian limits for residential and agricultural sectors which are40–50 mg/kg for each PHE and FLUT respectively [53]. The phe-nomenon was clearly observed for the MFP treatment shown inFig. 6 in which both PHE and FLUT residual concentrations werewithin the targeted limits after 24 h of reaction. Although SP hasbeen demonstrated here to have positive effects on PAH removal ineither single MF treatment or persulphate enhanced MF treatment,further investigations of soil quality such as SOM oxidation, cata-lyst dissolution as well as the effect of treatment on microorganismsand other soil properties are necessary prior to scaling-up.

4. Conclusions

This work has demonstrated that chelate enhanced MF treat-

ment results in rapid PAH degradation which obeys pseudo-first-order kinetics (9.15 × 10−2 to 2.30 × 10−1 min−1 and 7.20 × 10−2

to 2.34 × 10−1 min−1 for PHE and FLUT respectively). The resultsshowed that SP was the most effective CA for PAH removal from

8 ineeri

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Venny et al. / Chemical Eng

ontaminated soil (up to 89.13%). Both SP and MA could enhanceMW PAH oxidation even at natural soil pH.

Persulphate addition into the MF treatment could facilitate PAHemoval of up to 95.41% and 92.25% for PHE and FLUT respectivelyver 24 h of MFP oxidation. This suggested the compatibility of per-ulphate in enhancing MF treatment coupled with SP. AlthoughP is proven to have promising potential in PAH removal fromoil matrix, further investigations into its environmental impactshould be performed.

cknowledgements

This work was supported by the Ministry of Science, Technol-gy and Innovation (MOSTI), Malaysia under the eScienceFund3-02-12-SF0078. The Faculty of Engineering at the University ofottingham Malaysia Campus is also acknowledged for its support

owards this project.

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