Hydroxyl radicals based advanced oxidation processes (AOPs) for remediation of soils contaminated with organic compounds: a review

Accepted Manuscript

Review

Hydroxyl radicals based advanced oxidation processes (AOPs) for remediationof soils contaminated with organic compounds: a review

Min Cheng, Guangming Zeng, Danlian Huang, Cui Lai, Piao Xu, Chen Zhang,Yang Liu

PII: S1385-8947(15)01236-XDOI: http://dx.doi.org/10.1016/j.cej.2015.09.001Reference: CEJ 14136

To appear in: Chemical Engineering Journal

Received Date: 30 May 2015Revised Date: 1 September 2015Accepted Date: 2 September 2015

Please cite this article as: M. Cheng, G. Zeng, D. Huang, C. Lai, P. Xu, C. Zhang, Y. Liu, Hydroxyl radicals basedadvanced oxidation processes (AOPs) for remediation of soils contaminated with organic compounds: a review,Chemical Engineering Journal (2015), doi: http://dx.doi.org/10.1016/j.cej.2015.09.001

This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customerswe are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, andreview of the resulting proof before it is published in its final form. Please note that during the production processerrors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

1

Hydroxyl radicals based advanced oxidation processes (AOPs) for remediation of soils

contaminated with organic compounds：：：：a review

Min Cheng a,b, Guangming Zeng a,b,∗, Danlian Huang a,b,∗, Cui Lai a,b, Piao Xu a,b, Chen Zhang

a,b, Yang Liu a,b

a College of Environmental Science and Engineering, Hunan University, Changsha, Hunan

410082, China

b Key Laboratory of Environmental Biology and Pollution Control (Hunan University),

Ministry of Education, Changsha, Hunan 410082, China

∗ Corresponding author at: College of Environmental Science and Engineering, Hunan University, Changsha, Hunan 410082, China.

Tel.: +86–731– 88822754; fax: +86–731–88823701.

E-mail address: [email protected] (G.M. Zeng), [email protected] (D.L. Huang).

2

Abstract

Advanced oxidation processes (AOPs) constitute a promising technology for the

remediation of soils contaminated with non-easily removable organic compounds. This review

provides the reader with a general overview on the application of AOPs to pesticides,

polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs) and total

petroleum hydrocarbons (TPHs) contaminated soils remediation. Four types of AOPs

including Fenton processes, TiO2 photocatalysis, plasma oxidation and ozonation were

discussed. In particular, this paper focuses on the fundamental principles and governing

factors of the two typical techniques-Fenton oxidations and TiO2 photocatalysis. Apart from

the effect of chemical’s dosage as a major influencing factor, selected information such as

pollutant characteristics, light intensity, soil characteristics and pH are presented. Some

innovations (e.g., chelating agents, surfactants) on the traditional AOPs and the combined

utilization of AOPs with other techniques (e.g., bioremediation, soil washing) are also

documented and discussed. This review also highlights the effects of AOPs treatments on soil

properties.

Keywords: Advanced oxidation processes; Soil remediation; Fenton; Photocatalysis; Plasma

oxidation; Ozonation

1. Introduction

As an important component of the ecological environment, soil is one of the main

resources that human beings rely on to survive, and also the material repository of

bio-geochemical cycles. Nowadays, extensive use of pesticides and fertilizers constantly

damage farmland. In addition, accidental emissions of harmful pollutants, the industrial

wastewater and the landfill leachate have been causing serious soil pollution and deteriorating

soil quality [1]. Government and the public now have recognized the potential dangers that

organic pollutants such as pesticides, polycyclic aromatic hydrocarbons (PAHs),

polychlorinated biphenyls (PCBs) and total petroleum hydrocarbons (TPHs) posed to human

health and the environment [2-4]. Remediating the contaminated soil, in order to protect

human health and achieve sustainable development, has become the common view of both

government and public.

3

Extensive work has been devoted to the development of soil remediation techniques, and

several new and innovative solutions for efficient contaminants removal from soils have been

investigated to reduce the contaminants contents to a safe and acceptable level [5, 6]. Among

these treatment techniques, chemical oxidation has the potential for rapidly treating or

pretreating soils contaminated with toxic and biorefractory organic compounds [7]. Chemical

oxidation aims to mineralize the pollutants to carbon dioxide (CO2), water (H2O) and

inorganics or, at least, transform them into harmless or biodegradable products [8]. In last two

decades a lot of researches have been addressed to this aim and pointed out the prominent role

of a special class of oxidation techniques defined as advanced oxidation processes (AOPs),

which usually operated at or near ambient temperature and pressure [9, 10].

The advantage of AOPs over all chemical and biological processes is that they are totally

“environmental-friendly” as they neither transfer pollutants from one phase to the other (as in

chemical precipitation and adsorption) nor produce massive amounts of hazardous sludge [10].

AOPs are capable of degrading nearly all types of organic contaminants into harmless

products [11] and almost all rely on the production of reactive hydroxyl radicals (•OH) with a

redox potential of 2.8 V [12]. •OH is the second most reactive species next to fluorine atom,

they attack the most part of organic pollutants molecules with rate constants usually in the

order of 106-109 M−1 s−1, which is 106 -1012 times faster than ozone [13, 14]. In these process,

•OH initiate a series of oxidation reactions then leading to the ultimate mineralization

products of CO2 and H2O [15]. Due to these characteristics, numerous works have been done

to investigate the applications of AOPs to treat different types of contaminated soils. However,

they have concentrated solely on one technology employed in treating one kind of pollution,

and to date, an evaluation of all currently available AOPs for different types of contaminated

soils remediation has not been reported. With this in mind, the review attempts to summarize

and discuss the state of the art in the treatment of pesticides, PAHs, PCBs and TPHs

contaminated soils by using •OH based AOPs.

2. Advanced oxidation processes

The versatility of AOPs is also enhanced by the fact that they offer different possible

processes for •OH generation, thus allowing a better compliance with the specific treatment

requirements. The following techniques are most often used in AOPs (Fig. 1): (i) Fenton

4

oxidations, (ii) photocatalysis, (iii) plasma oxidation, and (iv) ozonation. As shown in Fig. 2,

the interest of researchers for AOPs began only around 1995 [12] and continues nowadays

since the number of investigations devoted to its application to soils remediation is still rising

considerably. Fig. 2 also demonstrates the majority of these researches were focused on

Fenton oxidations and photocatalysis, the publications about these two methods by 2015 are

466, which comprised 75.5 % of the total publications about AOPs (617).

2. 1. Fenton processes

As depicted in Fig. 3, Fenton reaction causes the decomposition of H2O2 and the

formation of highly reactive •OH (Eqs. (1)) [16] that can oxidize organic compounds (RH or

R) by hydrogen abstraction (R•) or by hydroxyl addition (•ROH). The highly reactive

molecules (R• and •ROH) can be further oxidized (Eqs. (2) and (3)) [17, 18]. Moreover, the

newly formed ferric ions (Fe3+) can catalyse H2O2 (Eqs. (4)), the reaction of H2O2 with Fe3+ is

known as a Fenton-like reaction [19]. Apart from Fe2+ regeneration, hydroperoxyl radicals

(HO2•) are produced in Fenton like reaction. HO2• are less sensitive than OH•, but they can

also attack organic contaminants [19].

Fe2+ + H2O2 → Fe3+ + HO− + •OH (1)

pH3, K3=70 M−1 s−1

RH + •OH→ H2O + R•→ further oxidations (2)

R + •OH→ •ROH → further oxidations (3)

Fe3+ + H2O2 �Fe2+ + H+ + HO2• (4)

K4=0.001-0.1 M−1s−1

Fenton/Fenton like processes can be carried out at room temperature and atmospheric

pressure; however, they are strongly dependent on the pH due to iron ions (Fe2+ and Fe3+) and

H2O2 speciation factors [20]. The optimum pH value for Fenton reaction is around 3, a slight

decrease or increase in the pH value will sharply reduce the efficiency of the systems [21, 22].

When pH goes below 3.0, H2O2 can solvate protons to form oxonium ions (H3O2+), which

would enhance the stability of H2O2 and reduce its reactivity with ferrous ions. And when pH

goes higher, the dissolved fraction of iron species would decrease as colloidal ferric species

appear [7]. To overcome this limitation, Fenton reaction has been modified to extend its range

of applicability to native soil pH which is at approximately neutral condition. Modified

5

Fenton reagents using chelating agents have particularly drawn much interest. Common

chelators include catechol, cyclodextrin, nitrilotriacetic acid and ethylenediaminetetraacetic

acid [23, 24]

Previous studies indicated that the degradation rate of organic pollutants by Fenton

oxidations could be strongly facilitated by ultraviolet visible (UV-Vis) light irradiation at

wavelength values higher than 300 nm [25, 26]. In this case, the photolysis of Fe3+ complexes

(Fe(OH)2+) allows Fe2+ regeneration and the occurrence of Fenton reaction due to the presence

of H2O2 [27]. This combined process (Eqs. (5) and (6)) is referred to as the photo-Fenton (or

photo-assisted Fenton) reaction [28]. In addition to the above reactions the formation of •OH

also occurs by other reactions in photo-Fenton process (Eqs. (7)). One point should be noted

is that the application of photo-Fenton processes requires strict pH control. The photo-Fenton

reaction is optimized at pH 2.8 [29, 30] where approximately half of the Fe(III) is present as

Fe3+ ion and half as Fe(OH)2+ ion. When pH goes above 2.8, the Fe(III) will precipitate as

oxyhydroxides and as pH goes lower the concentration of Fe(OH)2+ will decline [31, 32].

Fe2+ + H2O2 → Fe(OH)2+ + •OH (5)

Fe(OH)2++ hv → Fe2+ + •OH (6) H2O2

+ hv → 2•OH (7)

There is a great interest in electro-Fenton soil remediation [33, 34]. In electro-Fenton, a

direct current of low intensity is applied across electrode pairs implanted in the ground on

each side of the contaminated soil. In contrast to the classical Fenton process, H2O2 is

generated in situ at the cathode with O2 or air feeding [35]. The pollutants are destroyed by the

action of Fenton’s reagent in the bulk together with •OH generated at anode surface as shown

in Eqs. (8) [36]:

M + H2O → M (•OH) + H+ + e- (8)

where M represents the anode material. Electro-Fenton process is also considered as a clean

treatment for the soil washing solutions without any production of sludge [37, 38].

Furthermore, no iron would be needed, since the iron could be directly extracted from soil

[37]. In some cases, modified Fenton treatment is favored over conventional Fenton treatment.

For example, surfactants modified Fenton process can enhance the solubility of pollutants by

decreasing the interfacial tension, allowing the Fenton oxidation to treat aged soils [39]. The

6

detailed information about modified Fenton processes can be found in another review article

[40]. This article discussed both the mechanisms and application of Fenton processes for

environment purification.

2. 2. Photocatalysis

Photocatalysis, which makes use of the semiconductor metal oxide as catalyst is an

extensively studied field in the last four decades [41-45]. Many semiconductors, such as TiO2,

ZnO, CdS, GaP, WO3, and NiO, have been tested as photocatalysts [44-47]. TiO2 in the

anatase form was proved to be the most appropriate one due to its characteristics such as high

photoactivity, chemical inertness, non-toxic, low cost and easy to obtain [48]. On the contrary,

some other semiconductors, including ZnO, CdS and GaP, cannot be used for environmental

purification, because they can dissolve and produce toxic byproducts during the

photocatalysis of semiconductors [48].

Semiconductor molecules contain a valence band which occupied with stable energy

electrons and empty higher energy conduction band [49]. The band gap of TiO2 (anatase) is

3.2 eV, wavelength is about 400 nm [50, 51]. Many case studies on different substrates have

been successfully demonstrated the decomposition of organic pollutants in soils using TiO2

under UV-irradiation or solar light [52-55]. The initiating procedure of the photocatalytic

reaction is the absorption of the radiation with the formation of holes (h+) in valence band and

electrons (e−) in conduction band in femtosecond timescale (Eqs. (9)) [51].

TiO2 + hv → e- + h+ (9)

When appropriate scavengers (H2O and/or HO−) are present, oxidation reactions can take

place to form reactive •OH (Eqs. (10) and (11)) [56]. During the photocatalytic process, some

other reactive radicals like superoxide radical anion (O2•-) are also formed (Eqs. (12)) [57].

According to Eqs. (13)-(15), O2•- can lead to the production of •OH [64]. Besides, e− may also

react with some adsorbed contaminants through reductive processes directly (Eqs. (16)) [58].

Fig. 4 presents the schematic on removal of pollutants by the formation of photoinduced

charge carriers in the TiO2 particle surfaces.

TiO2 (h+) + H2O → TiO2 + •OH + H+ (10)

TiO2 (h+) + HO− → TiO2 + •OH (11)

O2 + e- → O2•

- (12)

7

O2•- + H+

→ HO2• (13)

HO2• + H+ + TiO2 (e−) → H2O2 + TiO2 (14)

H2O2 + TiO2 (e−) → HO• + HO− + TiO2 (15)

RH + h+ → •R + H+ (16)

The bandgap (3.2 eV) of conventional TiO2 requires UV light irradiation to activate the

photo-reaction. Whereas only approximately 4-5% of light reaching the Earth’s surface is UV

light and about 45% is visible light [42]. Hence, the energy requirement severely hinders the

treatment if the sun light is to be used as the energy source. To promote the energy efficiencies

of this treatment, a lot of efforts have been made. Previous studies have successfully extended

the photoresponse region of TiO2 to visible light by introducing additional components into

the lattice structure [59, 60]. Both non-metal (e.g., N, F, C, S) doping and metal (e.g., Cr, Co,

V, Fe) doping of TiO2 has shown great prospect in achieving visible light activated

photocatalysis [44, 60]. Among them, N-doped TiO2 seems to be the most efficient and

extensively investigated [59]. The interested readers will find more detailed information about

the fundamental aspects of visible light active TiO2 photocatalysts in other relevant and

excellent review papers [42, 46]. TiO2 photocatalysis have been fully developed in numbers of

case studies on water and air purification, and has been recognized as one of the most

promising environmental remediation technologies [42]. However, the photodegradation of

organic pollutants in soils is more complex because it may affected by many factors, such as

light absorption characteristics, humic substances content and moisture content.

2. 3. Plasma oxidation and ozonation

Plasmas oxidation is also regarded as highly competitive technology for the removal of

organic pollutants from soils. Plasmas oxidation was examined as an eco-innovative method

of soil remediation only in recent years. Currently, low-temperature plasma (LTPs), especially

the techniques based on pulsed corona discharge (PCD) [61, 62] and dielectric barrier

discharge (DBD) [63, 64] has received a great attention in soil remediation field. During the

plasma production, high energy electrons are generated, providing space charge and highly

reactive species, such as O, OH, and H radicals, and O3, H2O2 molecules [65]. In discharge

plasma processes, H2O2 is considered to be one of the most major active species involved in

the degradation of organic contaminants [66]. H2O2 can react with various organic

8

contaminants via direct oxidation or indirect oxidation. Generally, the indirect oxidation plays

a more important role due to •OH generation by H2O2 decomposition [66]. In discharge

plasma process, O3 can react with H2O2 to form •OH (Eqs. (17)), and high energy electrons

can attack H2O2 to generate •OH at the same time (Eqs. (18)) [66, 67]. Furthermore, •OH can

be produced via self-decomposition of H2O2 (Eqs. (7)) and Fenton processes [66].

H2O2 + O3 → •OH+ HO- + O2 (17)

H2O2 + e- → •OH+ OH- (18)

Amongst the technologies that can be applied “in situ” or “on site”, soil O3 application is

catalogued as one of the most promising systems. One method of O3 treatment is to cause the

organic contaminants decomposed directly by O3, and the other is to make them react

indirectly with •OH, which can be generated by the O3 decomposition [8]. Ozone

decomposition at pH > 6 is theorized to follow the indirect reaction pathway to form •OH

radicals [68]. Ozone has been shown to degrade pesticides, hydrocarbons, PAHs, and PCBs in

soils [69-71]. O3 could decompose on soil active surfaces (i.e. metal oxides, soil organic

matter, etc.) to generate •OH according to Eqs. (19) [72].

O3 + Soil → Soil-HO• + O2 (19)

3. AOPs for remediation of pesticides contaminated soils

Soil contamination by pesticides is a widespread occurrence [73]. Pesticides have been

used to mitigate or repel pests such as insects, bacteria, nematodes, mites and other organisms

that affect food production or human health since Second World War [74], and many times

irresponsible use has made them an environmental problem [75]. This is mainly due to their

properties, such as hard-biodegraded and high retention time in soil [76]. Furthermore, some

pesticides, such as dichlorodiphenyltrichloroethane (DDT), can exhibit phenomena known as

biomagnification [77], which means their concentration increases as they pass up the food

chain. Such compounds in soil can pose a constant threat to both humans and wildlife. In

order to reduce the risk of pesticides contaminated soils, a variety of methods have been

developed. AOPs are some of the techniques widely studied and applied to remove toxic and

biorefractory contaminants in soils. Summaries of some representative studies are compiled in

Table 2.

3.1 Fenton oxidation

9

Fenton processes are the most commonly used AOP in the remediation of pesticides

contaminated soils [78]. In an early study, Miller et al. [79] assessed the treatment of

pendimethalin-contaminated soil using Fenton reaction. Laboratory data revealed that nearly

all of the pendimethalin originally present in soil was removed by Fenton oxidation [79].

Subsequent studies have found that DDT, diuron, 2,4-dichlorophenol (2,4-DCP),

pentachlorophenol (PCP) and numbers of other pesticides in the soils can be effectively

degraded by Fenton/Fenton-like process [80-86]. Some works show that Fenton-like system

can achieve high pesticide removal at near-neutral pH by using chelates [85-86]. However,

there is also evidence that the direct application of the Fenton process is very aggressive to the

soil, and can be a disaster to the microbes in the soil [87, 88]. To overcome this limitation, the

coupled process (soil washing followed by Fenton oxidation) was investigated. Great progress

has been made in recent years, for example, about 95% of DDT was removed by Fenton

oxidation after soil washing using a Triton X-100 solution [89]. Nevertheless, it seemed that

the extraction solvents might impact on the removal efficacy to a large extent. In a similar

note, this combined process was used for the remediation of atrazine contaminated soils using

ethanol solution as extraction solvent, and it showed a degradation yield of only 28.1% [90].

This is believed to be a consequence of the nonselective oxidization nature of •OH, a higher

consumption of Fenton’s reagents was needed in the presence of ethanol. Apart from the

relatively lower removal rate, some attention must be given to the co-extraction of metals

present in the soil matrix, which may also be washed out during the soil washing process [89].

In such case, further treatment is needed, before reuse or safe discharge of the wastewater.

Several researchers have investigated the photo-Fenton reaction as a feasible treatment

method of pesticides contaminated soil. Huston et al. [32] observed that 13 pesticides in most

cases were completely removed after 30 minutes photo-Fenton oxidation. Study carried out by

Villa et al. [89] focused on the possibility of using photo-Fenton oxidation to degrade DDT

and dichlorodiphenyl dichloroethylene (DDE) in soil. They obtained the similar results that

most of DDT and DDE were decomposed after 6 hours treatment [89]. It should be noted that

a small amount of 2,2-bis (4-chlorophenyl)-1-chloroethylene (4,4’-DDMU) which is more

toxic than DDT, was generated as the intermediates of DDT degradation during the Fenton

treatment [86, 89]. Furthermore, the difference in efficacies between Fenton and photo-Fenton

10

processes for treating pesticides contaminated soil was studied. Laboratory data showed that

99% degradation for chlorimuron-ethyl was achieved in 10 minutes by using photo-Fenton

oxidation, however, only 68% degradation was observed after 30 minutes Fenton oxidation

[91]. These observations proved that UV radiation can remarkably improve the degradation

efficiency when the production of •OH in the absence of UV radiation was not sufficient to

reach a high level of mineralization [27]. It is reported that photo-Fenton process is able to use

solar radiation [91]. However, it should be noticed that these processes may not be suitable for

in situ treatment as the light cannot penetrate soil and therefore largely lowers the treatment

efficiency.

The influence of soil characteristics on Fenton treatment of pesticides contaminated soil

was also investigated. Several reports have suggested lower organic content and pH of the soil

lead to higher pesticides degradation ratio with Fenton oxidation [82, 92, 93]. This is mainly

due to the role of organic matter as a free radical scavenger, which would therefore compete

with the pesticides, and it was also affected by the mechanism of •OH formation which is

favored at lower pH values. Meanwhile, Fenton treatment changes soil properties. A

considerable amount of total organic carbon (TOC), chemical oxygen demand (COD),

biochemical oxygen demand (BOD), and nitrate may release during Fenton treatment. For

example, Miller [82] observed the content of TOC and nitrate increased by almost 10 times

after the treatment. This observation is consistent with a report by Gozzi et al. [91], they also

found that a periodic H2O2 addition was not able to improve the removal rate of TOC. On the

other hand, several researchers noted that photo-Fenton process could lead to higher TOC

removal rates. It was reported in presence of UV irradiation, only 29 mg L−1 DOC was

detected after the treatment [89], and in another study the removal ratio of TOC reached 95%

[91]. A proposed reason is in the photo-Fenton process, the recycling of the reaction Fe2+ to

Fe3+ by photolysis accelerates the production of •OH [94].

3.2 Photocatalytic degradation

Many works have been done try to establish TiO2 photocatalysis based technology for

the treatment of pesticides contaminated soil. One such attempt has been presented by

Higarashi and Jardim [95] who studied the remediation of diuron contaminated soil using

TiO2 with solar light irradiation. They observed that both TiO2 and the diuron content showed

11

no influence on the kinetics of diuron degradation [95]. Furthermore, they concluded that

when diuron concentration was within 100 mg kg−1, loading the soil with 0.1% of catalyst was

sufficient to achieve over 99% degradation [95]. This finding was confirmed by Xu et al. [96]

when studying the photocatalytic degradation of glyphosate (40 mg kg−1). In their study, the

best TiO2 loading amount was 0.5%, which is the lowest dosage used in their experiments.

The results suggest that only a small amount of catalyst was needed when the pollutants

content is low. And when the content of organic pollutants is high, a more catalyst load can

enhance the photodegradation significantly by producing more •OH [97-99].

On the other hand, the light intensity seems to be a key parameter of the

photodegradation. It was reported in the top 1 cm of soil the removal of diuron reached 90%,

however, when the depth comes to 8 cm, degradation of the pesticide decreased to only 5%

[95]. In a subsequent study [96], the removal rate of glyphosate was 87.3% with the soil

thickness of 0.22 cm, and then decreased to 38.21% when thickness reached 1.05 cm. A recent

study found that the optimum soil thickness for the photocatalytic oxidation of imidacloprid

was 0.2 cm [97]. All these data suggest that the photocatalytic reaction mainly occurs in the

surface part of soil and the degradation efficiency decreases as the soil layer becomes thicker.

The simplest explanation is the solar/UV light cannot penetrate into the soil, so the necessary

elements for the photocatalytic reaction in this part of soil are absent.

Humidity of the soil is another important parameter of the photodegradation. Higarashi

[95] found Diuron half-life dropped by 50% when the soil became more saturated with water.

Similar results have been observed in several other studies, for example, the most efficient

degradation of glyphosate in the laboratory was obtained when the moisture content was

between 30% and 50% [98]. These results suggest that the presence of appropriate amount of

water can promote the photodegradation, which is mainly because water formed suitable

conditions of transporting the pollutants to the catalyst surface, or increased the mobility of

the pollutants. Additionally, H2O molecules can further react in the surface of TiO2 by

oxidation in the hole (h+) and generate •OH, thereby increasing the removal efficiency [97].

Apart from humidity, some other soil characteristics (for example, humic substances) have

also been observed affect the photocatalytic degradation. As pointed out by Quan et al. [99],

the humic substances can inhibit the photodegradation by reducing the amount of light

12

available to excite the pesticides. Furthermore, it was found the sensitization of humic

substances can enhance of the photodegradation when the content of humic substances is low,

however, the abundance of humic substances might scavenge radicals and resulted in

inhibition of the photodegradation [100]. Some observations [99, 101-102] suggest that the

photodegradation rate increases with the pH, the mechanism underlying this stimulatory effect

is when the soil pH is higher, more •OH can be formed via photo-oxidation of OH−.

4. AOPs for remediation of PAHs contaminated soils

PAHs are ubiquitous environmental contaminants, which are by-products of fossil fuels

processing or combustion [103]. PAHs are included in the European Community and in the

Environmental Protection Agency priority pollutant list mainly due to their mutagenic and

carcinogenic properties [104-106]. Because of hydrophobic and recalcitrant characteristics,

PAHs tend to be adsorbed on solid particles and these characteristics make it as one of the

major soil pollutants [107]. Since most PAHs are non-volatile and hardly biodegradable,

conventional methods such as soil vapor extraction and bioventing cannot remove them

effectively [108]. With powerful oxidizing capacity, AOPs gradually reflect their advantages

in eliminating PAHs from contaminated soils [109-111]. Summaries of some representative

studies are compiled in Table 3.

4. 1. Fenton oxidation

Several reports have suggested that the removal efficiency of PAHs by direct Fenton

oxidation is affected by the characteristic of PAHs in a large extent. As for PAHs species, the

high molecular weight (HMW) PAHs (4-6 rings) are more lipophilic and less water soluble

than the lower molecular weight (LMW) PAHs (2-3 aromatic rings). In most cases, the

desorption of PAHs molecular from the soil particles is the rate limiting step, due to the

chemical oxidations were primarily take place in solution [112]. These properties determine

that the reactivity of different PAH species towards Fenton reaction, in general, the

degradation of the LMW PAHs was easier and faster than degradation of HMW PAHs. In one

laboratorial work, the degradation of 24 PAHs using Fenton oxidations was evaluated. Results

showed that 89 and 59% removal were achieved for PAHs with 2 and 3 rings, respectively,

whereas for PAHs with 4, 5, and 6 rings the corresponding figures varied between 0 and 38%

[112]. This study also shows that anthracene (3 rings), pyrene (4 rings) and benzo(a)pyrene (5

13

rings) were easier to be degraded than the other PAHs with similar structure [112]. Similar

trends have been observed in the other laboratories. Ferrarese et al. [113] found the removal

rates of 16 PAHs varied from 27 to 98% (80% removal for LMW PAHs and 57% removal for

HMW PAHs) in their experiments. In another case [114], 94.6% of phenanthrene (3 rings)

was removed after Fenton oxidation, but more than half of pyrene (4 rings) was still remained

in the system. Based on these observations, it can be concluded that HMW PAHs taking

longer to reach the solution thus decreases their susceptibility to Fenton oxidation. However,

there were some exceptions to this; for example, benzo(a)pyrene was more easily oxidized by

Fenton reaction than some smaller PAHs [112, 115]. This comes about because

benzo(a)pyrene has the higher ionisation potential than other PAHs with five rings and even

some smaller PAHs [115].

Studies have documented that modified Fenton oxidations can achieve better removal

efficiency and allow Fenton oxidations to treat more recalcitrant PAHs. In some cases, by

addition of chelating agents can prevent iron precipitation, and some reactive radicals may

form with •OH together. Venny et al. [116] evaluated the utilization of sodium pyrophosphate

as an inorganic chelating agent, and the results showed that the modified Fenton reaction

significantly enhanced the mineralization of PAHs. Different kind of chelating agents were

studied in the laboratory. Above 95% of total PAHs in the heavily contaminated soils was

degraded in another experimental work using catechol as the chelating agent [113]. There

might be a strong connection between Fenton oxidation efficiency and PAHs availability.

Several researchers suggested the major constraint of Fenton oxidations was caused by the

low availability of PAHs [117]. Surfactants were usually used to enhance the solubility of

PAHs by decreasing the interfacial tension and increasing their partitioning to the hydrophobic

cores of surfactant micelles. As a representative one, the cationic surfactant cetylpyridinium

chloride (CPC), was proved suitable for treating aged soils. It was reported that in the

presence of CPC, degradation ratio of pyrene increased from 91% (in the absence of CPC) to

97% in the pyrite Fenton system [117]. Besides, CPC would not cause secondary pollution,

because most of them can be degrade to CO2 and ammonium simultaneously during the

treatment [117]. Pre-heating (60-100 °C) followed by Fenton oxidation is a new approach

applied on the treatment of aged soils presenting low PAHs availability. It is suggested that the

14

pre-heating could induce a remobilization of PAHs through a change in the PAHs sorption

sites, thus enhance the oxidation efficiency [105]. As yet, there are only a limited number of

articles reported this method.

Several works have been done in order to acquire a precise understanding of the

influence of soil characteristics on the treatment. It has long been known that •OH may

consumed by TOC before they react with the target pollutants [115]. Bogan and Trbovic [118]

took a step further, they found that the susceptibility of PAHs to Fenton oxidation was a

function of TOC when above a threshold value of 5%, when the TOC is lower; the oxidation

is mainly depended on soil porosity. It was also revealed that HMW PAHs usually have

stronger affinities for humic acid, which makes them less susceptible to Fenton oxidations

[119]. On the other hand, the impacts of Fenton treatment on soil quality were investigated.

Gan et al. observed the decline in soil organic matter and C/N ratio after Fenton treatment

[120]. It is easy to understand that the some of the organic matter were oxidized together with

the contaminants due to the non-selective nature of •OH generated from Fenton processes.

The loss of N was not as much as C was due to the fact that some organic carbon was

converted to CO2 and escaped to the air [120]. Furthermore, it was pointed that the main

disadvantage of Fenton treatment is the reduction in soil pH, and for revegetation purpose,

Fenton treatment was appropriately adopted for soil with native pH >6.2 [120].

4.2. Photocatalytic degradation

Photocatalytic degradation is also an efficient method to decompose PAHs in the soils.

Using TiO2 under UV light to degrade PAHs has been studied in previous decade. In one of

the earliest studies [121], only 43.5% of pyrene on soil surfaces was removed under the

optimum conditions. To improve the degradation efficiency, Rababah and Matsuzawa [122]

developed a recirculating-type photocatalytic reactor. They observed that in the presence of

both TiO2 and H2O2 the degradation rate of fluoranthene was 99% compared to a relatively

lower degradation rate of 83% in the presence of TiO2 alone [122]. The enhancement in the

photocatalytic degradation rates by addition of H2O2 was confirmed by the later study [123].

Two mechanisms, indeed, can be responsible for the enhancement. Firstly, H2O2 could

enhance the degradation by providing additional •OH through trapping of photogenerated

electrons (e-), secondly, H2O2 self-decomposition by UV light would also produce •OH [123].

15

On the other hand, some works have been done on the modification of TiO2. They can be

summarized as morphological modifications. For example, increasing TiO2 particle surface

area can provide more reaction area thus lead to higher photocatalytic degradation efficiency.

Ireland et al. [124] studied the photocatalytic degradation of a mixture of 16 PAHs in a

contaminated soil using high surface area TiO2 as catalyst, and achieved high recoveries

(93-99%) of these PAHs. Some studies showed that catalytic technique using nanometer

(10-30 nm) anatase TiO2 could be attractive in remediation of persistent organic pollutants

contaminated soils [125]. To our surprise, no reports of the utilization of chemical modified

TiO2 for the remediation of PAHs contaminated soils was found. Considering the great

achievements that obtained in water and air remediation [49], the chemical modified TiO2

should have the same advantages in soil remediation. However, some experimental works are

needed to examine this hypothesis.

A number of efforts have been made to determine the main influences on the

photodegradation. Zhang et al. [126] investigated the effects of TiO2 load on photocatalytic

degradation of benzo(a)pyrene, pyrene and phenanthrene on soil surfaces. The results showed

that the variation of TiO2 dosage from 0.5% to 3% had no significant effect on the removal

rates. One possible explanation for this result is that even low TiO2 load can provide enough

catalyst surface area to promote maximum rates of destruction. In contrast, UV light intensity

was proved as a key factor in the process of photocatalytic degradation. Laboratory data

showed that the degradation increased with UV light intensity [127]. It was suggest that under

the higher light intensity, the electron-hole formation was predominant, thus the electron-hole

recombination was negligible [125]. Some researchers suggest that in the presence of humic

acid, the degradation of PAHs could be enhanced in a large extent [128]. It was demonstrated

that when humic substances absorbed UV irradiation, a variety of photochemical changes in

humic acids can lead to production of reactive oxygen species (e.g., singlet oxygen, peroxy

radicals) [125]. The reactive oxygen intermediates then attack PAHs thus facilitate their

degradation [129, 130]. The degradation rates were greatly affected by pH, under acidic or

alkaline conditions, more H+ or OH− ions were produced in soil, and these ions were able to

enhance the photocatalytic oxidation. For example, Fan et al. observed the highest pyrene and

benzo(a)pyrene photocatalytic degradation rates at acidic conditions (pH 4.2), and

16

phenanthrene was significantly photocatalytic degraded at alkaline conditions (pH 9.7) [131].

Besides, the photocatalytic degradation of PAHs was found promoted by increasing of soil

particle sizes and the processing temperature [123].

5. AOPs for remediation of PCBs contaminated soils

PCBs are toxic and persistent pollutants that have been used in a variety of applications

such as transformers, capacitors, coolants, and lubricants since 1930s [132]. Due to their

hazardous nature and chemical stability, they are categorized as persistent organic pollutants

[133]. Their extreme persistence in the environment and ability to bioconcentrate in the food

chain present a great environmental risk [134]. Although incineration and land-filling are

proven and widely used technologies for treating PCBs-contaminated soils, there is

widespread public opposition to these approaches [135]. Therefore, alternative remedial

technologies like AOPs are needed for PCBs destruction. Summaries of some representative

studies are compiled in Table 4.

5.1. Fenton oxidation

It is reported that the direct treatment of PCBs contaminated soils by employing the

Fenton process acquired remarkable efficiency [134, 136]. In a recently published work, 98%

removal of the original PCBs structure was obtained after 3 days [26]. The authors observed

that PCBs was mostly removed in the early period of the process. In the laboratory,

degradation rate up to 53% of PCBs was achieved in the first half hour, and about 94%

removal in 24 hours. This is probably affected by the •OH generation rate, which was found

decreasing with time [137]. The results indicate the direct Fenton oxidation is a

time-consuming process for the remediation of PCBs contaminated soils. In order to improve

the processing efficiency, photo-Fenton and modified Fenton oxidation were studied. The

results of bench-scale studies on photo-Fenton oxidation of PCBs with UV light (254 nm)

show that the removal rate of PCBs in 30 minutes increased to 98% from a value of 53%

obtained without UV light [26]. In another experiment, 100% removal of PCBs was achieved

using the photo-Fenton process, and no dangerous residues was detected [137]. The utilization

of catalyzed H2O2 propagations was also evaluated, based on traditional Fenton reaction, the

diluted H2O2 was slowly added to the system to generate •OH. Recent studies showed that two

PCBs-contaminated soils from Superfund sites were effectively treated using this process

17

[138].

Some researchers [138] have found the degree of degradation was also dependent on the

degree of chlorination, it was noted that the rate of oxidation increases as the content in

chlorine atoms decreases. The same phenomenon was observed when studying the oxidation

reaction of several chlorophenols with •OH in aqueous solutions [139]. This dependence on

the degree of chlorination is due to the fact that oxidation takes place through the addition of a

hydroxyl group to a non-halogenated position, thus generating species with greater reactivity

[140]. Apart from the steric blocking that the oxidation induced in the non-chlorinated

positions, the chlorinated positions are also non-reactive [137]. Lindsey et al. [136] showed

that in the presence of cyclodextrin, the degradation efficiency for PCBs was significant

increased. Cyclodextrin or derivatized cyclodextrin are capable of improving the efficiency of

Fenton treatment through simultaneous complexation of Fe2+ and PCBs. The removal

efficiency of PCBs doubled with the addition of cyclodextrin in the experiment [120]. Fenton

reaction generally needs low pH (usually around 3) to maintain iron ion solubility and prevent

the formation of iron hydroxides and oxides, nevertheless, utilize of stabilizers (chelating

agents) allows higher pH conditions [136, 139]. In such conditions, the chelating agents

chelate the iron, allowing the Fenton process operated at near neutral pH. For example, by

using cyclodextrin as the chelating agent, experiments conducted at natural PH (pH=6.3) gave

similar removal of PCBs with the experiment carried out at pH 3 without cyclodextrin [136].

5.2. Photocatalytic degradation

The study of photocatalytic degradation of PCBs in contaminated soil with TiO2 was

firstly reported by Chiarenzelli et al. [141]. They found direct photocatalytic treatment can be

a viable remediation technology for lesser chlorinated PCBs which are more active and

mobile, over 80% of them were eliminated after a 24 h irradiation without pretreatment or

amendments [141]. However, when highly chlorinated PCBs are present, the releasing of

PCBs from aged-contaminated soil will be the critical procedure to the photodegradation. To

conquer this problem, photocatalytic degradation with added surfactants was investigated

[142, 143]. In these processes, surfactants were used as solubilizing agents to desorb PCBs

from aged soils, makes it easy for the followed photocatalytic degradation. Results showed

PCBs in the aged soil can be effectively photodegraded [143]. As has been reported by several

18

authors, soil washing and subsequent TiO2 photocatalytic degradation is another viable

alternative [144-146]. For example, polyoxyethylene lauryl ether was proved to be a good

extraction solvent since it supported the feasibility of both soil washing and photocatalytic

degradation [144]. However, extraction solvents like cyclodextrins and Tween80 are

undesirable because they might improve the stability of PCBs and protect PCBs from

photocatalytic degradation [146].

Several researches have been carried out on optimization of photocatalytic remediation

of PCBs in contaminated soil. Laboratory data indicated that solar light intensity has little

impact on the photodegradation, but the efficiency of the degradation can be enhanced

significantly by using lower wavelength UV light. A proposed reason is UV light can increase

the solubility and accessibility of PCBs to the photocatalytic reactions through the

hydroxylation [147, 148]. The content of organic matter was found very important to the

photocatalytic oxidation [146]. It has long been known that organic matter in the soils may

inhibit the free radical chain reactions [146]. On the other hand, desorption of PCBs

molecules into the solution is more easy for soils with lower organic matter content due to the

fact that the organic matter contain a significant fraction of the adsorption sites. It was

observed that TiO2 load has no significant effect on the removal rates, and the removal rates

even decreased slightly when the TiO2 load was beyond its optimum range [149]. Similar

results was found by Zhu et al. [146] who reported the degradation rate drastically decreased

from 92% to 66% when the TiO2 content in the extract increased from 50 to 500 mg L−1. The

reduction in the removal rates is probably due to the scattering effect of the high concentration

of TiO2 on UV [150]. Zhou et al. observed the photodegradation rate of PCBs increased as the

pH value increased [145]. The mechanism is at high initial pH, more •OH can be formed via

photo-oxidation of OH−, and when pH was too low, •OH was difficult to form, especially when

considerable photon energy was blocked by TiO2 [145].

6. AOPs for remediation of TPHs contaminated soils

Nowadays petroleum hydrocarbon pollution has been one of the major environmental

problems, not only by their toxicity but also by the significant amounts released. Soil

contaminated with petroleum hydrocarbon is adverse to plant growth and is also a potential

source of groundwater pollution [151]. Bioremediation has proven to be successful in many

19

case studies [152]. However, bioremediation usually requires a long treatment period, and it is

often inefficient to lower the contamination level below the stringent environmental cleanup

standards [153]. From another point of view, AOPs are showing great potential as feasible

technology for remediation of TPHs contaminated soils [154]. Summaries of some

representative studies are compiled in Table 5.

6.1. Fenton oxidation

Among all the chemical technologies for remediation of TPHs contaminated soil, Fenton

oxidation can lead to the best yields in pollutants degradation [155], and also cheaper than

many other technologies for the clear up of TPHs contaminated soils [156]. Different kinds of

iron catalysts including iron (III) sulfate, iron (III) nitrate, iron (II) and iron (III) perchlorate

for the treatment have been investigated. Laboratory data showed that iron (III) nitrate and

iron (III) perchlorate were most effective in degrading TPHs in the soil [157, 158]. Tsai and

Kao [159] assessed the potential of applying Fenton-like oxidation using a special catalyst

(waste basic oxygen furnace slag) to treat the diesel contaminated soil. The experiments

showed a good performance result (96%) because of the basic oxygen furnace contains a

significant amount of extractable irons such as soluble iron and amorphous iron.

Some researchers made innovation on the traditional Fenton method, as a representative

one, ultrasonic energy was applied to facilitate the degradation. On the one hand, more

petroleum hydrocarbon are able to be desorbed from the soil due to the application of

ultrasound, on the other hand, ultrasonic energy can promote H2O2 decomposition in •OH

[160, 161]. The experimental results show that, 96 % of the original TPHs were eliminated

from the soil in the combined Fenton and ultrasound treatment, however, when in the absence

of ultrasound, removal rate decreased to 21% [162]. The feasibility of using iron electrode

corrosion to enhance the Fenton oxidation was evaluated. It was observed that

electrokinetic-Fenton oxidation resulted in higher TPHs removal efficiency (97%) compared

to the efficiency observed from Fenton oxidation (27%) alone [163]. The iron electrode can

supply iron continuously during the Fenton-like process, thus lead to higher TPHs removal

efficiency. The combination of soil washing and electro-Fenton process has been studied

recently. It was reported that the electro-Fenton oxidation achieved excellent mineralization of

the TPHs in the eluates. However, this technology needs to be improved, since the efficiency

20

of the washing process was very low [164]. The use of chelating agents or stabilizers was

scarcely studied for the treatment of TPHs contaminated soils, and only a limited number of

publications have appeared in the last 2 years. Traditional chelates like EDTA and trisodium

citrate can stabilize the oxidant, but low removal rate of TPHs was obtained because of both

the high pH achieved and the competition of pollutant and chelant for the oxidant [165]. Some

researchers suggest that SOM and fractions thereof (e.g., humic acid, fulvic acid) can be

considered as effective chelates, because they can increase the participation of native Fe

oxides in Fenton processes [166].

The combination of Fenton and biological remediation processes has become a popular

technology for treatment of TPHs contaminated soil in recent years. The traditional Fenton

processes were proved effective but also require a good amount of chemical agents. In some

situation, Fenton processes integrated with biological remediation can be more economic and

environmental friendly. Pre-treatment with Fenton oxidations can destroy petroleum

hydrocarbons and convert them into more biodegradable compounds, thus increase the

effectiveness of soil remediation and economic feasibility [167]. It was recorded that with

Fenton pre-treatment, petroleum oil removal rate of the biodegradation enhanced 52% [168].

However, it should be noted that too much H2O2 dose would retard the followed biological

treatment and lead to a poor pollutants removal efficiency. The mechanism responsible for the

inhibition is: exothermic Fenton’s reactions with high H2O2 concentration may result in a

sharp temperature increase which would destroy the native microbiota in the soil and even

cause undesired soil sterilization [87, 169]. In order to avoid the sharp temperature increase

when large amounts of H2O2 was added to the soil in one time, graded modified Fenton

process was employed, in which H2O2 was added intermittently. Normally, 3-time addition of

H2O2 was desirable and economical due to lower enhancement of TPHs removal between

3-time addition and more times addition [170].

6.2. Photocatalytic degradation

In recent years, there have been abundant publications focus on the application of TiO2

photocatalysis to the treatment of organic pollutants, to date, however, there is hardly any

information about the treatment on TPHs contaminants soils. Some researcher in Poland

evaluated photocatalysts based on TiO2 for the remediation of spent motor oil contaminated

21

soil in 1990s [171]. It is reported that even at the best conditions, the best degree of oil

decomposition was only 37.6% after 4 hours photocatalytic process [171]. The relatively low

degradation rate is probably because sun light cannot penetrate into the soil (2cm) and the

photocatalytic reaction occurs only on the soil surface. More work is required to investigate

the feasibility of photocatalytic degradation for the TPHs contaminated soil remediation.

7. Plasma oxidation and ozonation for remediation of contaminated soils

7.1. Plasma oxidation

The first study that uses LTPs discharges to treat polluted soil was carried out by Redolfi

et al. [172] who evaluated kerosene components oxidation in a soil matrix by a DBD reactor

at atmospheric pressure. Results showed that the total kerosene components abatement can

reaches 90%, and the removal mechanism was determined as the oxidation of kerosene in the

soil matrix [172]. In the sequent studies [62, 173, 174], Wang et al. studied the remediation of

PCP contaminated soil using PCD plasma, and the promising results were obtained. The

results indicated that PCP degradation efficiency increased with an increase in peak pulse

voltage or pulse frequency. This is due to the enhancement of energy input which contributes

to the increase in active species [174]. The pollution time showed small effect on PCP

degradation, whereas granular size of the soil was found very important [173]. It was ascribed

to the fact that the soil with smaller granular size would allow more contact area for active

species to react with organic compounds in soil [173, 175]. In addition, enhancing soil pH and

lowering humic acid in soil were found to be favorable for PCP degradation efficiency.

Alkaline condition is favorable for O3 to be decomposed into the powerful •OH [174], which

is also found in the other experiment [67]. It was observed that humic acid was partially

degraded during the discharge process [173]. That indicates the competitive reaction between

humic acid and PCP with active species, thereby resulting in the decrease of PCP degradation

efficiency. Lower p-nitrophenol (PNP) mineralization efficiency was observed in deeper soil

layers in another work reported by Wang et al. [176]. This is attributed to both the diffusion

behavior of the active species in soil layers and the degradation characteristics of PNP [176].

DBD plasma was also examined as a method for the ex situ remediation of non-aqueous

phase liquid (NAPL)-contaminated soils. The NAPL (100,000 mg kg-1) remediation

efficiency was found as high as 99.9% after 2 minutes of DBD plasma treatment in a

22

plane-to-grid reactor [63]. However, in a cylinder-to-plane DBD reactor, NAPL remediation

efficiency decreases as NAPL concentration increases from 1,000 to 100,000 mg kg-1 and high

energy densities are needed to achieve the high removal of NAPL [177]. It is suggested that

the most volatile NAPL compounds are evaporated and then oxidized in gas phase, whereas

the less volatile compounds are evaporated and oxidized in gas phase and soil matrix [177].

Conventionally, discharge plasma occurred in gas phase firstly, and then the generated

chemically active species permeated into contaminated soil layer to oxidize pollutants; in that

case, some short-lived active species would disappear before entering soil layer and

participating in pollutants degradation [61]. Recently, direct multi-channel pulsed discharge

plasma in soil was developed to remediate contaminated soil [61, 66]. In this approach, the

discharge plasma was triggered directly in contaminated soil, which can enhance the

utilization efficiency of chemically active species [66]. The experimental data show that

pollutant degradation efficiency and energy yield obtained by direct discharge plasma in soil

were comparable with those obtained by indirect discharge plasma out of soil [62]. In some

other works, the pulsed discharge plasma-TiO2 catalytic (PDPTC) technique was investigated

to enhance the remediation efficiency [178, 179]. Experimental results showed that 88.8% of

PNP could be removed in 10 min in the PDPTC system, compared with 78.1% in plasma

alone system [179]. Compared with plasma alone system, the enhancement effect on PNP

mineralization is attributed to more amounts of chemically active species (e.g., O3 and H2O2)

produced in the PDPTC system [178].

7.2. Ozonation

The need for effective in situ treatment technologies has led to the increased use of ozone

to remediate the contaminated soil. Experiments have found that soil ozonation and

contaminant removal efficiency are affected by both the soil properties (moisture, structure,

pH, etc.) and chemical properties of the contaminants. Soil with larger pore spaces can

provide better transport of ozone through the soil matrices [180]. Studies have suggested that

the removal efficiency of pollutants by ozone decreases in the presence of soil moisture

compared to that observed in dry soils [181, 182]. In one of these studies, pyrene removal

reached 94.9% in dry soils compared to 55.5% and 33.8% removal in 5% and 10% moisture

soils, respectively [182]. This is because soil moisture reduces the number of active sites

23

where ozone can react with the pollutants on the soil particles [180, 182]. The transport of

gas-phase ozone can be significantly retarded by O3 consumption due to reactions with SOM

[71]. Furthermore, it is reported that the order of the reactivity of the fractions of SOM is:

aromatic > aliphatic > polar [4]. The experimental results also show that the treatment

efficiency of ozonation increased with the increase of soil pH from 2 to 8 [182]. This is

because the indirect reactions with •OH (O3 decomposition at pH > 6) can facilitate the

pollutant oxidation [68, 182]. However, the opposite results were observed when the pH value

increased to 12 [70]. The decreased removal at high pH may be due to loss of reactive species

by free-radical scavenging [70].

The degradation of LMW PAHs was found more efficient than that of HMW PAHs. For

example, Masten and Davies [183] reported 95% removal of phenanthrene, 91% removal of

pyrene while chrysene was reduced to only 50%. It is because HMW PAHs normally have

strong bond-localization energies and a high affinity for soil organic matter. PCBs and PAHs

in soils contaminated in a long-term are more strongly bound to the soil sorption complex

leading to lower removal efficiencies compared to freshly contaminated soils [184, 185].

Studies showed that the more strongly adsorbed contaminants would require higher ozone

dosages for removal, whereas gas flow-rate does not affect the process efficiency [72, 184].

Luster-Teasley et al. [182] reported that increasing the pH of the soil from 2.0 to 8.0 obtained

a 141.6% enhancement in pyrene removal. The proposed reason is the higher pH is favorable

for •OH generation [182]. On the other hand, attention should be paid to the formation of

carboxylic acid during the ozonation which decreased the soil pH to 3.0 from an initial value

of 6.0 [186]. This can result in soil acidification which will restrict plant respiration and

increase the metal mobility potential such as lead (Pb) in the environment.

Hong et al. [187] developed an ozonation technique that incorporated rapid, successive

cycles of pressurization (690 kPa) and depressurization, and this technique was more effective

than conventional ozonation treatment. Near complete or complete removal of PCBs and

PAHs was achieved in 30 minutes. This efficiency was due to soil aggregate fracturing upon

pressure cycles that exposed the contaminants, as well as by the confluence of PCB and O3 at

the gas–liquid interface in the presence of microbubbles [187].

8. Conclusions and prospects

24

This review provides the reader with a general overview on the treatments of pesticides

and PAHs, PCBs and TPHs contaminated soils by using AOPs, with a special address to the

two mainly applied methods Fenton processes and photocatalytic processes. Fenton process

has become popular because of the following reasons: (1) easy to implement, (2) able to

degrade a wide range of contaminants, (3) sub-products are usually harmless or biodegradable.

The main drawback of conventional Fenton treatment is the reduction in soil pH. And large

quantity of oxidant is needed for the soil with high content of organic matter or the additional

substances (e.g., chelating agents, surfactants). Besides, the delivery of oxidants to the

contaminated zones is difficult because H2O2 can be decomposed by iron oxides and enzymes

(e.g., catalases and peroxidases) presenting in the soil. Recently, the application of

photocatalytic processes has been extended to treatment of contaminated soils. TiO2

photocatalyst was widely studied owing to its characteristics such as safety, high

photocatalytic activity and low cost. The major defect of TiO2 photocatalysis is that the

photocatalytic degradation only occurs in the soil surface and removal efficiency decreases as

the soil layer becomes thicker. Meanwhile, the lack of visible light activity also hinders its

practical applications. And using UV-lamp can be costly due to the limited lamp life. Plasma

oxidation can almost completely remove the pollutants from soils in minutes. And plasma

oxidation is able to treat soils with high concentration pollutants. However, some short-lived

active species would disappear before entering soil layer and participating in pollutants

degradation, and high energy densities are needed to treat the heavily polluted soil. Ozone

oxidation process also has a rapid treatment time and high degradation efficiency. Gaseous

ozone is advantageous over aqueous oxidants such as Fenton’s reagents because of its

relatively easier delivery to unsaturated porous media. But, it seems that ozonation is only

suitable for treating soils with low moisture content.

Conventional Fenton process was found effective to treat soils with lower organic

content and pH. Some innovations on the traditional Fenton method have been made, for

example, UV light was applied to assist the degradation. In some situation, Fenton method is

often integrated with biological remediation which is more economic and environmental

friendly. Laboratory data indicated that the rate and level of photocatalytic degradation was

enhanced by the employ of lower wavelength UV light and under acidic conditions. Normally,

25

loading the contaminated soil with 0.5 wt% of TiO2 was sufficient to achieve good removal

efficiency, yet more TiO2 may need when the contaminant content is high. Enhancing soil pH

and lowering humic acid content were found to be favorable for plasma oxidation. Ozonation

is suitable to treat soils with large pore spaces and low moisture.

Consider the drawback that its optimal pH is around 3, the traditional Fenton treatment

will become less popular. More work needs to be done on modified Fenton treatments with

stabilizers or chelating agents. For example, solve the problem that chelating agents can

compete for the •OH with pollutants and result in a significant loss of efficiency. Fenton

processes integrated with biological remediation could become a widespread application

because it’s both low-cost and environmental friendly. Although the field application of soil

extraction to enhance the treatment efficiency seems difficult to implement, it will continue to

be a good approach for the treatment of aged soils which presenting low contaminants

availability. The utilization of sun light is currently limited by the photo-inefficiency of the

TiO2. In order to be photo-excited under visible light and aim at solar-driven TiO2

photocatalysis, some strategies are needed to modify the TiO2 catalyst. Future studies will

probably focus on the visible light activated photocatalysis catalyzed by morphological or

chemical modified TiO2. The study on plasma oxidation started in the recent years, and has

obtained promising results. More work is needed to be done to put this technique into

practical use. Ozonation can be good alternative because of its high degradation efficiency

and potential for in situ treatments.

Acknowledgements

This study was financially supported by the National Natural Science Foundation of

China (51378190, 51278176, 51408206), the Environmental Protection Technology Research

Program of Hunan (2007185), the Fundamental Research Funds for the Central Universities,

the Hunan University Fund for Multidisciplinary Developing (531107040762), the Program

for New Century Excellent Talents in University (NCET-13-0186), the Program for

Changjiang Scholars and Innovative Research Team in University (IRT-13R17) and a Project

Supported by Scientific Research Fund of Hunan Provincial Education Department

(521293050).

References

[1] M. Manz, K.D. Wenzel, U. Dietze, G. Schüürmann, Persistent organic pollutants in

26

agricultural soils of central Germany, Sci. Total Environ. 277 (2001) 187-198.

[2] Z. Zhang, J. Jiatieli, D. Liu, F. Yu, S. Xue, W. Gao, Y. Li, D.D. Dionysiou, Microwave

induced degradation of parathion in the presence of supported anatase- and rutile-TiO2/AC

and comparison of their catalytic activity, Chem. Eng. J. 231 (2013) 84-93.

[3] G. Zeng, M. Chen, Z. Zeng, Risks of neonicotinoid pesticides, Science, 340 (2013)

1403-1403.

[4] M. Gómez-Alvarez, T. Poznyak, E. Ríos-Leal, C. Silva-Sánchez, Anthracene

decomposition in soils by conventional ozonation, J. Environ. Manage. 113 (2012) 545-551.

[5] R.D. Villa, A.G. Trovó, R.F.P. Nogueira, Environmental implications of soil remediation

using the Fenton process, Chemosphere 71 (2008) 43-50.

[6] X. He, A.A. de la Cruz, K.E. O'Shea, D.D. Dionysiou, Kinetics and mechanisms of

cylindrospermopsin destruction by sulfate radical-based advanced oxidation processes, Water.

Res. 63 (2014) 168-178.

[7] R.J. Watts, P.C. Stanton, J. Howsawkeng, A.L. Teel, Mineralization of a sorbed polycyclic

aromatic hydrocarbon in two soils using catalyzed hydrogen peroxide, Water. Res. 36 (2002)

4283-4292.

[8] R. Andreozzi, V. Caprio, A. Insola, R. Marotta, Advanced oxidation processes (AOP) for

water purification and recovery, Catal. Today 53 (1999) 51-59.

[9] J.A. Khan, X. He, H.M. Khan, N.S. Shah, D.D. Dionysiou, Oxidative degradation of

atrazine in aqueous solution by UV/H2O2/Fe2+, UV//Fe2+ and UV//Fe2+ processes: A

comparative study, Chem. Eng. J. 218 (2013) 376-383.

[10] K. Ayoub, E.D. van Hullebusch, M. Cassir, A. Bermond, Application of advanced

oxidation processes for TNT removal: A review, J. Hazard. Mater. 178 (2010) 10-28.

[11] S. Esplugas, P.L. Yue, M.I. Pervez, Degradation of 4-chlorophenol by photolytic

oxidation, Water. Res. 28 (1994) 1323-1328.

[12] S.J. Masten, S.H. Davies, The use of ozonation to degrade organic contaminants in

wastewaters, Environ. Sci. Technol. 28 (1994) 180A-185A.

[13] J. Hoigné, H. Bader, Rate constants of reactions of ozone with organic and inorganic

compounds in water—II: dissociating organic compounds, Water. Res. 17 (1983) 185-194.

[14] Y. Zhang, C. Han, G. Zhang, D.D. Dionysiou, M.N. Nadagouda, PEG-assisted synthesis

27

of crystal TiO2 nanowires with high specific surface area for enhanced photocatalytic

degradation of atrazine, Chem. Eng. J. 268 (2015) 170-179.

[15] R.C. Weast, Handbook of chemistry and physics, The American Journal of the Medical

Sciences, 257 (1969) 423.

[16] F. Haber, J. Weiss, The catalytic decomposition of hydrogen peroxide by iron salts, in:

Proceedings of the Royal Society of London A: Mathematical, Physical and Engineering

Sciences, The Royal Society, 1934, pp. 332-351.

[17] V. Flotron, C. Delteil, Y. Padellec, V. Camel, Removal of sorbed polycyclic aromatic

hydrocarbons from soil, sludge and sediment samples using the Fenton’s reagent process,

Chemosphere 59 (2005) 1427-1437.

[18] N. Kang, I. Hua, P.S.C. Rao, Enhanced Fenton’s destruction of non-aqueous phase

perchloroethylene in soil systems, Chemosphere 63 (2006) 1685-1698.

[19] J. De Laat, H. Gallard, Catalytic decomposition of hydrogen peroxide by Fe (III) in

homogeneous aqueous solution: mechanism and kinetic modeling, Environ. Sci. Technol. 33

(1999) 2726-2732.

[20] A.A. Burbano, D.D. Dionysiou, M.T. Suidan, T.L. Richardson, Oxidation kinetics and

effect of pH on the degradation of MTBE with Fenton reagent, Water. Res. 39 (2005)

107-118.

[21] Venny, S. Gan, H.K. Ng, Inorganic chelated modified-Fenton treatment of polycyclic

aromatic hydrocarbon (PAH)-contaminated soils, Chem. Eng. J. 180 (2012) 1-8.

[22] R. Matta, K. Hanna, T. Kone, S. Chiron, Oxidation of 2,4,6-trinitrotoluene in the

presence of different iron-bearing minerals at neutral pH, Chem. Eng. J. 144 (2008) 453-458.

[23] F.J. Rivas, Polycyclic aromatic hydrocarbons sorbed on soils: a short review of chemical

oxidation based treatments, J. Hazard. Mater. 138 (2006) 234-251.

[24] C.L. Yap, S. Gan, H.K. Ng, Fenton based remediation of polycyclic aromatic

hydrocarbons-contaminated soils, Chemosphere 83 (2011) 1414-1430.

[25] Y. Wang, C.S. Liu, F.B. Li, C.P. Liu, J.B. Liang, Photodegradation of polycyclic

aromatic hydrocarbon pyrene by iron oxide in solid phase, J. Hazard. Mater. 162 (2009)

716-723.

[26] J. Quiroga, A. Riaza, M. Manzano, Chemical degradation of PCB in the contaminated

28

soils slurry: Direct Fenton oxidation and desorption combined with the photo-Fenton process,

J. Environ. Sci. Heal. Part A 44 (2009) 1120-1126.

[27] R.G. Zepp, B.C. Faust, J. Hoigne, Hydroxyl radical formation in aqueous reactions (pH

3-8) of iron (II) with hydrogen peroxide: the photo-Fenton reaction, Environ. Sci. Technol. 26

(1992) 313-319.

[28] G. Ruppert, R. Bauer, G. Heisler, The photo-Fenton reaction—an effective

photochemical wastewater treatment process, J. Photochem. Photobiol. A: Chem. 73 (1993)

75-78.

[29] J.J. Pignatello, Dark and photoassisted iron (3+)-catalyzed degradation of chlorophenoxy

herbicides by hydrogen peroxide, Environ. Sci. Technol. 26 (1992) 944-951.

[30] A. Khataee, F. Salahpour, M. Fathinia, B. Seyyedi, B. Vahid, Iron rich laterite soil with

mesoporous structure for heterogeneous Fenton-like degradation of an azo dye under visible

light, J. Ind. Eng. Chem. 26 (2015) 129-135.

[31] V. Kavitha, K. Palanivelu, The role of ferrous ion in Fenton and photo-Fenton processes

for the degradation of phenol, Chemosphere 55 (2004) 1235-1243.

[32] P.L. Huston, J.J. Pignatello, Degradation of selected pesticide active ingredients and

commercial formulations in water by the photo-assisted Fenton reaction, Water. Res. 33

(1999) 1238-1246.

[33] M. Rodrigo, N. Oturan, M. Oturan, Electrochemically assisted remediation of pesticides

in soils and water: a review, Chem. Rev. 114 (2014) 8720-8745.

[34] I. Sirés, E. Brillas, M.A. Oturan, M.A. Rodrigo, M. Panizza, Electrochemical advanced

oxidation processes: today and tomorrow. A review, Environ. Sci. Pollut. Res.21 (2014)

8336-8367.

[35] E. Mousset, L. Frunzo, G. Esposito, E.D. van Hullebusch, N. Oturan, M.A. Oturan, A

complete phenol oxidation pathway obtained during electro-Fenton treatment and validated by

a kinetic model study, Appl. Catal. B: Environ. 180 (2016) 189-198.

[36] E. Mousset, N. Oturan, E.D. van Hullebusch, G. Guibaud, G. Esposito, M.A. Oturan,

Treatment of synthetic soil washing solutions containing phenanthrene and cyclodextrin by

electro-oxidation. Influence of anode materials on toxicity removal and biodegradability

enhancement, Appl. Catal. B: Environ. 160 (2014) 666-675.

29

[37] E. Mousset, M.A. Oturan, E.D. Van Hullebusch, G. Guibaud, G. Esposito, Soil

washing/flushing treatments of organic pollutants enhanced by cyclodextrins and integrated

treatments: state of the art, Crit. Rev. Env. Sci. Technol. 44 (2014) 705-795.

[38] C.A. Martínez-Huitle, E. Brillas, Decontamination of wastewaters containing synthetic

organic dyes by electrochemical methods: a general review, Appl. Catal. B: Environ. 87

(2009) 105-145.

[39] K. Nam, W. Rodriguez, J.J. Kukor, Enhanced degradation of polycyclic aromatic

hydrocarbons by biodegradation combined with a modified Fenton reaction, Chemosphere 45

(2001) 11-20.

[40] J.J. Pignatello, E. Oliveros, A. MacKay, Advanced oxidation processes for organic

contaminant destruction based on the Fenton reaction and related chemistry, Crit. Rev.

Environ. Sci. Technol. 36 (2006) 1-84.

[41] X. Chen, S.S. Mao, Titanium dioxide nanomaterials: synthesis, properties, modifications,

and applications, Chem. Rev. 107 (2007) 2891-2959.

[42] M. Pelaez, N.T. Nolan, S.C. Pillai, M.K. Seery, P. Falaras, A.G. Kontos, P.S. Dunlop,

J.W. Hamilton, J.A. Byrne, K. O'Shea, M.H. Entezari, D.D. Dionysiou, A review on the

visible light active titanium dioxide photocatalysts for environmental applications, Appl.

Catal. B: Environ. 125 (2012) 331-349.

[43] W. Liao, T. Zheng, P. Wang, S. Tu, W. Pan, Efficient microwave-assisted photocatalytic

degradation of endocrine disruptor dimethyl phthalate over composite catalyst ZrOx/ZnO, J.

Environ. Sci. 22 (2010) 1800-1806.

[44] M. Iwasaki, M. Hara, H. Kawada, H. Tada, S. Ito, Cobalt ion-doped TiO2 photocatalyst

response to visible light, J. Colloid. Interf. Sci. 224 (2000) 202-204.

[45] J. Yu, X. Yu, Hydrothermal synthesis and photocatalytic activity of zinc oxide hollow

spheres, Environ. Sci. Technol. 42 (2008) 4902-4907.

[46] H. Park, Y. Park, W. Kim, W. Choi, Surface modification of TiO2 photocatalyst for

environmental applications, J. Photochem. Photobiol. C: Photochem. Rev. 15 (2013) 1-20.

[47] A. Chen, G. Zeng, G. Chen, X. Hu, M. Yan, S. Guan, C. Shang, L. Lu, Z. Zou, G. Xie,

Novel thiourea-modified magnetic ion-imprinted chitosan/TiO2 composite for simultaneous

removal of cadmium and 2, 4-dichlorophenol, Chem. Eng. J. 191 (2012) 85-94.

30

[48] R. Daghrir, P. Drogui, D. Robert, Modified TiO2 for environmental photocatalytic

applications: a review, Ind. Eng. Chem. Res. 52 (2013) 3581-3599.

[49] A. Mills, S. Le Hunte, An overview of semiconductor photocatalysis, J. Photochem.

Photobiol. A: Chem. 108 (1997) 1-35.

[50] K. Nakata, A. Fujishima, TiO2 photocatalysis: Design and applications, J. Photochem.

Photobiol. C: Photochem. Rev. 13 (2012) 169-189.

[51] A. Fujishima, X. Zhang, D.A. Tryk, TiO2 photocatalysis and related surface phenomena,

Surf. Sci. Rep. 63 (2008) 515-582.

[52] T. Ochiai, A. Fujishima, Photoelectrochemical properties of TiO2 photocatalyst and its

applications for environmental purification, J. Photochem. Photobiol. C: Photochem. Rev. 13

(2012) 247-262.

[53] S.G. Kumar, L.G. Devi, Review on modified TiO2 photocatalysis under UV/visible light:

selected results and related mechanisms on interfacial charge carrier transfer dynamics, J.

Phys. Chem. A 115 (2011) 13211-13241.

[54] D. Fabbri, A. Crime, M. Davezza, C. Medana, C. Baiocchi, A.B. Prevot, E. Pramauro,

Surfactant-assisted removal of swep residues from soil and photocatalytic treatment of the

washing wastes, Appl. Catal. B: Environ. 92 (2009) 318-325.

[55] E. Pelizzetti, C. Minero, V. Carlin, E. Borgarello, Photocatalytic soil decontamination,

Chemosphere 25 (1992) 343-351.

[56] M.R. Hoffmann, S.T. Martin, W. Choi, D.W. Bahnemann, Environmental applications of

semiconductor photocatalysis, Chem. Rev. 95 (1995) 69-96.

[57] A.L. Linsebigler, G. Lu, J.T. Yates Jr, Photocatalysis on TiO2 surfaces: principles,

mechanisms, and selected results, Chem. Rev. 95 (1995) 735-758.

[58] M.A. Fox, M.T. Dulay, Heterogeneous photocatalysis, Chem. Rev. 93 (1993) 341-357.

[59] Y. Zhang, C. Han, M.N. Nadagouda, D.D. Dionysiou, The fabrication of innovative

single crystal N,F-codoped titanium dioxide nanowires with enhanced photocatalytic activity

for degradation of atrazine, Appl. Catal. B: Environ. 168–169 (2015) 550-558.

[60] O. Sacco, V. Vaiano, C. Han, D. Sannino, D.D. Dionysiou, Photocatalytic removal of

atrazine using N-doped TiO2 supported on phosphors, Appl. Catal. B: Environ. 164 (2015)

462-474.

31

[61] T.C. Wang, G. Qu, J. Li, D. Liang, Evaluation of the potential of soil remediation by

direct multi-channel pulsed corona discharge in soil, J. Hazard. Mater. 264 (2014) 169-175.

[62] T.C. Wang, G. Qu, J. Li, D. Liang, Remediation of p-nitrophenol and pentachlorophenol

mixtures contaminated soil using pulsed corona discharge plasma, Sep. Purif. Technol. 122

(2014) 17-23.

[63] C. Aggelopoulos, P. Svarnas, M. Klapa, C. Tsakiroglou, Dielectric barrier discharge

plasma used as a means for the remediation of soils contaminated by non-aqueous phase

liquids, Chem. Eng. J. 270 (2015) 428-436.

[64] S. Ognier, J. Rojo, Y. Liu, X. Duten, S. Cavadias, L. Thannberger, Mechanisms of

Pyrene Degradation during Soil Treatment in a Dielectric Barrier Discharge Reactor, Plasma

Processes Polym. 11 (2014) 734-744.

[65] J. Lou, N. Lu, J. Li, T. Wang, Y. Wu, Remediation of chloramphenicol-contaminated soil

by atmospheric pressure dielectric barrier discharge, Chem. Eng. J. 180 (2012) 99-105.

[76] T. Wang, G. Qu, Q. Sun, D. Liang, S. Hu, Formation and roles of hydrogen peroxide

during soil remediation by direct multi-channel pulsed corona discharge in soil, Sep. Purif.

Technol. 147 (2015) 17-23.

[67] T.C. Wang, G. Qu, J. Li, N. Lu, Transport characteristics of gas phase ozone in soil

during soil remediation by pulsed discharge plasma, Vacuum 101 (2014) 86-91.

[68] J. Hoigne, H. Bader, W.R. Haag, J. Staehelin, Rate constants of reactions of ozone with

organic and inorganic compounds in water - III. Inorganic compounds and radicals, Water

Res. 19 (1985) 993–1004.

[69] D.Y. Yu, N. Kang, W. Bae, M.K. Banks, Characteristics in oxidative degradation by

ozone for saturated hydrocarbons in soil contaminated with diesel fuel, Chemosphere 66

(2007) 799-807.

[70] A.C. Pierpoint, C.J. Hapeman, A. Torrents, Ozone treatment of soil contaminated with

aniline and trifluralin, Chemosphere 50 (2003) 1025-1034.

[71] J. Kim, H. Choi, Modeling in situ ozonation for the remediation of nonvolatile

PAH-contaminated unsaturated soils, J. Contam. Hydrol. 55 (2002) 261-285

[72] J. Rivas, O. Gimeno, G. Ruth, F.J. Beltrán, Ozone treatment of PAH contaminated soils:

operating variables effect, J. Hazard. Mater. 169 (2009) 509-515.

32

[73] S.M. Dasappa, R.C. Loehr, Toxicity reduction in contaminated soil bioremediation

processes, Water. Res. 25 (1991) 1121-1130.

[74] H.M. van der Werf, Assessing the impact of pesticides on the environment, Agr. Ecosyst.

Environ. 60 (1996) 81-96.

[75] G. Zeng, Z. Yu, Y. Chen, J. Zhang, H. Li, M. Yu, M. Zhao, Response of compost

maturity and microbial community composition to pentachlorophenol (PCP)-contaminated

soil during composting, Bioresour. Technol. 102 (2011) 5905-5911.

[76] G.K. Frampton, S. Jänsch, J.J. Scott‐Fordsmand, J. Römbke, P.J. Van den Brink,

Effects of pesticides on soil invertebrates in laboratory studies: a review and analysis using

species sensitivity distributions, Environ. Toxicol. Chem. 25 (2006) 2480-2489.

[77] T. Castelo-Grande, P.A. Augusto, D. Barbosa, Removal of pesticides from soil by

supercritical extraction—a preliminary study, Chem. Eng. J. 111 (2005) 167-171.

[78] E.C. Catalkaya, F. Kargi, Advanced oxidation and mineralization of simazine using

Fenton's reagent, J. Hazard. Mater. 168 (2009) 688-694.

[79] C.M. Miller, R.L. Valentine, M.E. Roehl, P.J. Alvarez, Chemical and microbiological

assessment of pendimethalin-contaminated soil after treatment with Fenton's reagent, Water.

Res. 30 (1996) 2579-2586.

[80] R. Andreozzi, I. Di Somma, R. Marotta, G. Pinto, A. Pollio, D. Spasiano, Oxidation of 2,

4-dichlorophenol and 3, 4-dichlorophenol by means of Fe (III)-homogeneous photocatalysis

and algal toxicity assessment of the treated solutions, Water. Res. 45 (2011) 2038-2048.

[81] E. Bocos, C. Fernández-Costas, M. Pazos, M.Á. Sanromán, Removal of PAHs and

pesticides from polluted soils by enhanced electrokinetic-Fenton treatment, Chemosphere 125

(2015) 168-174.

[82] R.M. Liou, S.-H. Chen, M.Y. Hung, C.S. Hsu, Catalytic oxidation of pentachlorophenol

in contaminated soil suspensions by Fe+3-resin/H2O2, Chemosphere 55 (2004) 1271-1280.

[83] M. Cao, L. Wang, L. Wang, J. Chen, X. Lu, Remediation of DDTs contaminated soil in a

novel Fenton-like system with zero-valent iron, Chemosphere 90 (2013) 2303-2308.

[84] H. Zhou, Q. Sun, X. Wang, L. Wang, J. Chen, J. Zhang, X. Lu, Removal of 2,

4-dichlorophenol from contaminated soil by a heterogeneous ZVI/EDTA/Air Fenton-like

system, Sep. Purif. Technol. (2014).

33

[85] J.M. Rosas, F. Vicente, E.G. Saguillo, A. Santos, A. Romero, Remediation of soil polluted

with herbicides by Fenton-like reaction: Kinetic model of diuron degradation, Appl. Catal. B:

Environ. 144 (2014) 252-260.

[86] F. Vicente, A. Santos, E.G. Sagüillo, Á.M. Martínez-Villacorta, J.M. Rosas, A. Romero,

Diuron abatement in contaminated soil using Fenton-like process, Chem. Eng. J. 183 (2012)

357-364.

[87] S.H. Ferguson, A.Z. Woinarski, I. Snape, C.E. Morris, A.T. Revill, A field trial of in situ

chemical oxidation to remediate long-term diesel contaminated Antarctic soil, Cold Reg. Sci.

Technol. 40 (2004) 47-60.

[88] M.J. Monahan, A.L. Teel, R.J. Watts, Displacement of five metals sorbed on kaolinite

during treatment with modified Fenton's reagent, Water. Res. 39 (2005) 2955-2963.

[89] R.D. Villa, A.G. Trovó, R.F.P. Nogueira, Soil remediation using a coupled process: soil

washing with surfactant followed by photo-Fenton oxidation, J. Hazard. Mater. 174 (2010)

770-775.

[90] L. Di Palma, P. Ferrantelli, E. Petrucci, Experimental study of the remediation of atrazine

contaminated soils through soil extraction and subsequent peroxidation, J. Hazard. Mater. 99

(2003) 265-276.

[91] F. Gozzi, A. Machulek Jr, V.S. Ferreira, M.E. Osugi, A.P.F. Santos, J.A. Nogueira, R.F.

Dantas, S. Esplugas, S.C. de Oliveira, Investigation of chlorimuron-ethyl degradation by

Fenton, photo-Fenton and ozonation processes, Chem. Eng. J. 210 (2012) 444-450.

[92] B.W. Tyre, R.J. Watts, G.C. Miller, Treatment of four biorefractory contaminants in soils

using catalyzed hydrogen peroxide, J. Environ. Qual. 20 (1991) 832-838.

[93] R.J. Watts, M.D. Udell, P.A. Rauch, S.W. Leung, Treatment of

pentachlorophenol-contaminated soils using Fenton's reagent, Hazard. Waste. Hazard. Mater.

7 (1990) 335-345.

[94] L.A. Pérez-Estrada, S. Malato, W. Gernjak, A. Agüera, E.M. Thurman, I. Ferrer, A.R.

Fernández-Alba, Photo-Fenton degradation of diclofenac: identification of main intermediates

and degradation pathway, Environ. Sci. Technol. 39 (2005) 8300-8306.

[95] M.M. Higarashi, W.F. Jardim, Remediation of pesticide contaminated soil using TiO2

mediated by solar light, Catal. Today 76 (2002) 201-207.

34

[96] X. Xu, F. Ji, Z. Fan, L. He, Degradation of glyphosate in soil photocatalyzed by

Fe3O4/SiO2/TiO2 under solar light, Inter. J. Env. Res. Pub. Heal. 8 (2011) 1258-1270.

[97] T. Sharma, A.P. Toor, A. Rajor, Photocatalytic degradation of imidacloprid in soil:

application of response surface methodology for the optimization of parameters, RSC Adv. 5

(2015) 25059-25065.

[98] M. Muneer, C. Boxall, Photocatalyzed degradation of a pesticide derivative glyphosate in

aqueous suspensions of titanium dioxide, Int. J. Photoenergy. 2008 (2008).

[99] X. Quan, X. Zhao, S. Chen, H. Zhao, J. Chen, Y. Zhao, Enhancement of p, p'-DDT

photodegradation on soil surfaces using TiO2 induced by UV-light, Chemosphere (2005).

[100]K.H. Chan, W. Chu, Effect of humic acid on the photolysis of the pesticide atrazine in a

surfactant-aided soil-washing system in acidic condition, Water. Res. 39 (2005) 2154-2166.

[101] X. Zhao, X. Quan, S. Chen, H. Zhao, Y. Liu, Photocatalytic remediation of

gamma-hexachlorocyclohexane contaminated soils using TiO2 and montmorillonite composite

photocatalyst, J. Environ. Sci-China 19 (2006) 358-361.

[102] L. Sanchez, J. Peral, X. Domenech, Photocatalyzed destruction of aniline in

UV-illuminated aqueous TiO2 suspensions, Electrochim. Acta 42 (1997) 1877-1882.

[103] C.A. Menzie, B.B. Potocki, J. Santodonato, Exposure to carcinogenic PAHs in the

environment, Environ. Sci. Technol. 26 (1992) 1278-1284.

[104] L. Samsøe-Petersen, E.H. Larsen, P.B. Larsen, P. Bruun, Uptake of trace elements and

PAHs by fruit and vegetables from contaminated soils, Environ. Sci. Technol. 36 (2002)

3057-3063.

[105] C. Biache, C. Lorgeoux, S. Andriatsihoarana, S. Colombano, P. Faure, Effect of

pre-heating on the chemical oxidation efficiency: Implications for the PAH availability

measurement in contaminated soils, J. Hazard. Mater. 286 (2015) 55-63.

[106] A.A. Dadkhah, A. Akgerman, Hot water extraction with in situ wet oxidation: PAHs

removal from soil, J. Hazard. Mater. 93 (2002) 307-320.

[107] A. Haritash, C. Kaushik, Biodegradation aspects of polycyclic aromatic hydrocarbons

(PAHs): a review, J. Hazard. Mater. 169 (2009) 1-15.

[108] K.J. Doick, N.M. Dew, K.T. Semple, Linking catabolism to cyclodextrin extractability:

determination of the microbial availability of PAHs in soil, Environ. Sci. Technol. 39 (2005)

35

8858-8864.

[109] H.-H. Liste, M. Alexander, Butanol extraction to predict bioavailability of PAHs in soil,

Chemosphere 46 (2002) 1011-1017.

[110] C. Cuypers, T. Grotenhuis, J. Joziasse, W. Rulkens, Rapid persulfate oxidation predicts

PAH bioavailability in soils and sediments, Environ. Sci. Technol. 34 (2000) 2057-2063.

[111] M.M. Shah, D.P. Barr, N. Chung, S.D. Aust, Use of white rot fungi in the degradation

of environmental chemicals, Toxicol. Lett. 64 (1992) 493-501.

[112] S. Jonsson, Y. Persson, S. Frankki, B. van Bavel, S. Lundstedt, P. Haglund, M.

Tysklind, Degradation of polycyclic aromatic hydrocarbons (PAHs) in contaminated soils by

Fenton's reagent: A multivariate evaluation of the importance of soil characteristics and PAH

properties, J. Hazard. Mater. 149 (2007) 86-96.

[113] E. Ferrarese, G. Andreottola, I.A. Oprea, Remediation of PAH-contaminated sediments

by chemical oxidation, J. Hazard. Mater. 152 (2008) 128-139.

[114] V.L.d. Silva, B.d.B. Neto, M.O. Simonnot, Phenanthrene and pyrene oxidation in

contaminated soils using Fenton's reagent, J. Hazard. Mater. 161 (2009) 967-973.

[115] S. Lundstedt, Y. Persson, L. Öberg, Transformation of PAHs during ethanol-Fenton

treatment of an aged gasworks’ soil, Chemosphere 65 (2006) 1288-1294.

[116] Venny, S. Gan, H.K. Ng, Modified Fenton oxidation of polycyclic aromatic

hydrocarbon (PAH)-contaminated soils and the potential of bioremediation as post-treatment,

Sci. Total Environ. 419 (2012) 240-249.

[117] K. Choi, S. Bae, W. Lee, Degradation of pyrene in cetylpyridinium chloride-aided soil

washing wastewater by pyrite Fenton reaction, Chem. Eng. J. 249 (2014) 34-41.

[118] B.W. Bogan, V. Trbovic, Effect of sequestration on PAH degradability with Fenton’s

reagent: roles of total organic carbon, humin, and soil porosity, J. Hazard. Mater. 100 (2003)

285-300.

[119] B.W. Bogan, V. Trbovic, J.R. Paterek, Inclusion of vegetable oils in Fenton’s chemistry

for remediation of PAH-contaminated soils, Chemosphere 50 (2003) 15-21.

[120] S. Gan, C.L. Yap, H.K. Ng, Investigation of the impacts of ethyl lactate based Fenton

treatment on soil quality for polycyclic aromatic hydrocarbons (PAHs)-contaminated soils, J.

Hazard. Mater. 262 (2013) 691-700.

36

[121] D. Dong, P. Li, X. Li, Q. Zhao, Y. Zhang, C. Jia, P. Li, Investigation on the

photocatalytic degradation of pyrene on soil surfaces using nanometer anatase TiO2 under UV

irradiation, J. Hazard. Mater. 174 (2010) 859.

[122] A. Rababah, S. Matsuzawa, Treatment system for solid matrix contaminated with

fluoranthene. II––Recirculating photodegradation technique, Chemosphere 46 (2002)

49-57.

[123] D. Dong, P. Li, X. Li, C. Xu, D. Gong, Y. Zhang, Q. Zhao, P. Li, Photocatalytic

degradation of phenanthrene and pyrene on soil surfaces in the presence of nanometer rutile

TiO2 under UV-irradiation, Chem. Eng. J. 158 (2010) 378-383.

[124] J.C. Ireland, B. Davila, H. Moreno, S.K. Fink, S. Tassos, Heterogeneous photocatalytic

decomposition of polyaromatic hydrocarbons over titanium dioxide, Chemosphere 30 (1995)

965-984.

[125] J. Gu, D. Dong, L. Kong, Y. Zheng, X. Li, Photocatalytic degradation of phenanthrene

on soil surfaces in the presence of nanometer anatase TiO2 under UV-light, J. Environ.

Sci-China 24 (2012) 2122.

[126] L. Zhang, P. Li, Z. Gong, X. Li, Photocatalytic degradation of polycyclic aromatic

hydrocarbons on soil surfaces using TiO2 under UV light, J. Hazard. Mater. 158 (2008)

478-484.

[127] C.S. Chang, C. Chang, S. Chen, M. Wang, R.M. Madhava, V.S. Satya, Effect of

sunlight irradiation on photocatalytic pyrene degradation in contaminated soils by micro-nano

size TiO2, Sci. Total Environ. 409 (2011) 4101.

[128] L. Zhang, C. Xu, Z. Chen, X. Li, P. Li, Photodegradation of pyrene on soil surfaces

under UV light irradiation, J. Hazard. Mater. 173 (2010) 168-172.

[129] T. Mill, Predicting photoreaction rates in surface waters, Chemosphere 38 (1999)

1379-1390.

[130] I.K. Konstantinou, A.K. Zarkadis, T.A. Albanis, Photodegradation of selected

herbicides in various natural waters and soils under environmental conditions, J. Environ.

Qual. 30 (2001) 121-130.

[131] F. Xiaozhen, L. Bo, G. Aijunb, Dynamics of solar light photodegradation behavior of

atrazine on soil surface, J. Hazard. Mater. 117 (2005) 75-79.

37

[132] M. Robson, S. Harrad, Chiral PCB signatures in air and soil: Implications for

atmospheric source apportionment, Environ. Sci. Technol. 38 (2004) 1662-1666.

[133] T. Harner, D. Mackay, K.C. Jones, Model of the long-term exchange of PCBs between

soil and the atmosphere in the southern UK, Environ. Sci. Technol. 29 (1995) 1200-1209.

[134] M. Manzano, J. Perales, D. Sales, J. Quiroga, Catalyzed hydrogen peroxide treatment of

polychlorinated biphenyl contaminated sandy soils, Water Air Soil Pollut. 154 (2004) 57-69.

[135] J.E. Duffy, M.A. Anderson, C.G. Hill, W.A. Zeltner, Wet peroxide oxidation of

sediments contaminated with PCBs, Environ. Sci. Technol. 34 (2000) 3199-3204.

[136] M.E. Lindsey, G. Xu, J. Lu, M.A. Tarr, Enhanced Fenton degradation of hydrophobic

organics by simultaneous iron and pollutant complexation with cyclodextrins, Sci. Total

Environ. 307 (2003) 215-229.

[137] A. Riaza-Frutos, J. Quiroga, M. Manzano, Remediation of contaminated soils with

PCBs using an integrated treatment: Desorption and oxidation, J. Environ. Eng. 133 (2007)

541-547.

[138] M. Ahmad, M. Simon, A. Sherrin, M. Tuccillo, J. Ullman, A. Teel, R. Watts, Treatment

of polychlorinated biphenyls in two surface soils using catalyzed H₂O₂ propagations,

Chemosphere 84 (2011) 855.

[139] N. Oturan, M. Panizza, M.A. Oturan, Cold incineration of chlorophenols in aqueous

solution by advanced electrochemical process electro-Fenton. Effect of number and position

of chlorine atoms on the degradation kinetics, J. Phys. Chem. A 113 (2009) 10988-10993.

[140] C.Y. Kuo, S.L. Lo, Oxidation of aqueous chlorobiphenyls with photo-Fenton process,

Chemosphere 38 (1999) 2041-2051.

[141] J. Chiarenzelli, R. Scrudato, M. Wunderlich, D. Rafferty, K. Jensen, G. Oenga, R.

Roberts, J. Pagano, Photodecomposition of PCBs absorbed on sediment and industrial waste:

implications for photocatalytic treatment of contaminated solids, Chemosphere 31 (1995)

3259-3272.

[142] W. Chu, K. Chan, C. Kwan, C. Jafvert, Acceleration and quenching of the photolysis of

PCB in the presence of surfactant and humic materials, Environ. Sci. Technol. 39 (2005)

9211-9216.

[143] Q. Huang, C.S. Hong, TiO2 photocatalytic degradation of PCBs in soil-water systems

38

containing fluoro surfactant, Chemosphere 41 (2000) 871-879.

[144] X. Zhu, D. Zhou, Y. Wang, L. Cang, G. Fang, J. Fan, Remediation of polychlorinated

biphenyl-contaminated soil by soil washing and subsequent TiO2 photocatalytic degradation,

J. Soils Sediments 12 (2012) 1371-1379.

[145] Z. Zhou, Y. Zhang, H. Wang, T. Chen, W. Lu, The comparative photodegradation

activities of pentachlorophenol (PCP) and polychlorinated Biphenyls (PCBs) using UV alone

and TiO2-derived photocatalysts in methanol soil washing solution, PloS one 9 (2014)

e108765.

[146] X. Zhu, D. Zhou, L. Cang, Y. Wang, TiO2 photocatalytic degradation of

4-chlorobiphenyl as affected by solvents and surfactants, J. Soils Sediments 12 (2012)

376-385.

[147] H.L. Liu, Y.R. Chiou, Optimal decolorization efficiency of Reactive Red 239 by

UV/TiO2 photocatalytic process coupled with response surface methodology, Chem. Eng. J.

112 (2005) 173-179.

[148] I.J. Ochuma, R.P. Fishwick, J. Wood, J.M. Winterbottom, Photocatalytic oxidation of 2,

4, 6-trichlorophenol in water using a cocurrent downflow contactor reactor (CDCR), J.

Hazard. Mater. 144 (2007) 627-633.

[149] K. Wong, S. Tao, R. Dawson, P. Wong, Optimization of photocatalytic oxidation of 2, 2

′, 3, 3′-tetrachlorobiphenyl, J. Hazard. Mater. 109 (2004) 149-155.

[150] R. ALBERICI, W. JARDIM, Photocatalytic degradation of phenol and chlorinated

phenols using Ag-TiO2 in a slurry reactor, Water. Res. 28 (1994) 1845-1849.

[151] M.H. Huesemann, Predictive model for estimating the extent of petroleum hydrocarbon

biodegradation in contaminated soils, Environ. Sci. Technol. 29 (1995) 7-18.

[152] A. Murugesan, C. Umarani, R. Subramanian, N. Nedunchezhian, Bio-diesel as an

alternative fuel for diesel engines—A review, Renewable Sustainable Energy Rev. 13 (2009)

653-662.

[153] G. Płaza, G. Nałęcz-Jawecki, K. Ulfig, R.L. Brigmon, The application of bioassays as

indicators of petroleum-contaminated soil remediation, Chemosphere 59 (2005) 289-296.

[154] M. Lu, Z. Zhang, W. Qiao, Y. Guan, M. Xiao, C. Peng, Removal of residual

contaminants in petroleum-contaminated soil by Fenton-like oxidation, J. Hazard. Mater. 179

39

(2010) 604-611.

[155] L. Mater, R. Sperb, L. Madureira, A. Rosin, A. Correa, C. Radetski, Proposal of a

sequential treatment methodology for the safe reuse of oil sludge-contaminated soil, J.

Hazard. Mater. 136 (2006) 967-971.

[156] C. Shi, R. Spence, Designing of cement-based formula for solidification/stabilization of

hazardous, radioactive, and mixed wastes, Crit. Rev. Environ. Sci. Technol. 34 (2004)

391-417.

[157] S.H. Kong, R.J. Watts, J.H. Choi, Treatment of petroleum-contaminated soils using iron

mineral catalyzed hydrogen peroxide, Chemosphere 37 (1998) 1473-1482.

[158] R.J. Watts, S.E. Dilly, Evaluation of iron catalysts for the Fenton-like remediation of

diesel-contaminated soils, J. Hazard. Mater. 51 (1996) 209-224.

[159] T. Tsai, C. Kao, Treatment of petroleum-hydrocarbon contaminated soils using

hydrogen peroxide oxidation catalyzed by waste basic oxygen furnace slag, J. Hazard. Mater.

170 (2009) 466-472.

[160] H. Zhang, J. Zhang, C. Zhang, F. Liu, D. Zhang, Degradation of CI Acid Orange 7 by

the advanced Fenton process in combination with ultrasonic irradiation, Ultrason. Sonochem.

16 (2009) 325-330.

[161] J.H. Sun, S.P. Sun, J.Y. Sun, R.X. Sun, L.P. Qiao, H.Q. Guo, M.H. Fan, Degradation of

azo dye Acid black 1 using low concentration iron of Fenton process facilitated by ultrasonic

irradiation, Ultrason. Sonochem. 14 (2007) 761-766.

[162] R. Flores, G. Blass, V. Dominguez, Soil remediation by an advanced oxidative method

assisted with ultrasonic energy, J. Hazard. Mater. 140 (2007) 399-402.

[163] T.T. Tsai, J. Sah, C.M. Kao, Application of iron electrode corrosion enhanced

electrokinetic-Fenton oxidation to remediate diesel contaminated soils: a laboratory feasibility

study, Journal of hydrology, 380 (2010) 4-13.

[164] D. Huguenot, E. Mousset, E.D. van Hullebusch, M.A. Oturan, Combination of

surfactant enhanced soil washing and electro-Fenton process for the treatment of soils

contaminated by petroleum hydrocarbons, J. Environ. Manage. 153 (2015) 40-47.

[165] F. Pardo, J.M. Rosas, A. Santos, A. Romero, Remediation of a biodiesel

blend-contaminated soil by using a modified Fenton process, Environ. Sci. Pollut. Res. 21

40

(2014) 12198-12207.

[166] M.K. Sherwood, D.P. Cassidy, Modified Fenton oxidation of diesel fuel in arctic soils

rich in organic matter and iron, Chemosphere 113 (2014) 56-61.

[167] X.B. Gong, Remediation of weathered petroleum oil-contaminated soil using a

combination of biostimulation and modified Fenton oxidation, Int. Biodeterior. Biodegrad. 70

(2012) 89-95.

[168] A. Goi, N. Kulik, M. Trapido, Combined chemical and biological treatment of oil

contaminated soil, Chemosphere 63 (2006) 1754-1763.

[169] C. Valderrama, R. Alessandri, T. Aunola, J. Cortina, X. Gamisans, T. Tuhkanen,

Oxidation by Fenton's reagent combined with biological treatment applied to a

creosote-comtaminated soil, J. Hazard. Mater. 166 (2009) 594-602.

[170] J. Xu, L. Xin, T. Huang, K. Chang, Enhanced bioremediation of oil contaminated soil

by graded modified Fenton oxidation, J. Environ. Sci. 23 (2011) 1873-1879.

[171] M. Hamerski, J. Grzechulska, A. Morawski, Photocatalytic purification of soil

contaminated with oil using modified TiO2 powders, Sol. Energy 66 (1999) 395-399.

[172] M. Redolfi, C. Makhloufi, S. Ognier, S. Cavadias, Oxidation of kerosene components in

a soil matrix by a dielectric barrier discharge reactor, Process Saf. Environ. 88 (2010)

207-212.

[173] T.C. Wang, N. Lu, J. Li, Y. Wu, Evaluation of the potential of pentachlorophenol

degradation in soil by pulsed corona discharge plasma from soil characteristics, Environ. Sci.

Technol. 44 (2010) 3105–3110.

[174] T.C. Wang, N. Lu, J. Li, Y. Wu, Degradation of pentachlorophenol in soil by pulsed

corona discharge plasma, J. Hazard. Mater. 180 (2010) 436–441.

[175] T.C. Wang, G. Qu, J. Li, N. Lu, Kinetics studies on pentachlorophenol degradation in

soil during pulsed discharge plasma process, J. Electrostat. 71 (2013), 994-998.

[176] T.C. Wang, G. Qu, J. Li, D. Liang, S. Hu, Depth dependence of p-nitrophenol removal

in soil by pulsed discharge plasma, Chem. Eng. J. 239 (2014) 178-184.

[177] C.A. Aggelopoulos, C.D. Tsakiroglou, S. Ognier, S. Cavadias, Non-aqueous phase

liquid-contaminated soil remediation by ex situ dielectric barrier discharge plasma, Int. J.

Environ. Sci. Technol. 12 (2014) 1011-1022.

41

[178] T.C. Wang, N. Lu, J. Li, Y. Wu, Plasma-TiO2 Catalytic Method for High-Efficiency

Remediation of p-Nitrophenol Contaminated Soil in Pulsed Discharge, Environ. Sci. Technol.

45 (2011) 9301-9307.

[179] T.C. Wang, N. Lu, J. Li, Y. Wu, Y. Su. Enhanced degradation of p-nitrophenol in soil in

a pulsed discharge plasma-catalytic system, J. Hazard. Mater. 195 (2011) 276-280.

[180] M.M. O’Mahony, A.D. Dobson, J.D. Barnes, I. Singleton, The use of ozone in the

remediation of polycyclic aromatic hydrocarbon contaminated soil, Chemosphere 63 (2006)

307-314.

[181] H. Zhang, L. Ji, F. Wu, J. Tan, In situ ozonation of anthracene in unsaturated porous

media, J. Hazard. Mater. 120 (2005) 143-148.

[182] S. Luster-Teasley, N. Ubaka-Blackmoore, S. Masten, Evaluation of soil pH and

moisture content on in-situ ozonation of pyrene in soils, J. Hazard. Mater. 167 (2009)

701-706.

[183] S.J. Masten, S.H. Davies, Efficacy of in-situ for the remediation of PAH contaminated

soils, J. Contam. Hydrol. 28 (1997) 327-335.

[184] H. Javorská, P. Tlustoš, M. Komárek, D. Leštan, R. Kaliszová, J. Száková, Effect of

ozonation on polychlorinated biphenyl degradation and on soil physico-chemical properties, J.

Hazard. Mater. 161 (2009) 1202-1207.

[185] J.J. Pignatello, B. Xing, Mechanisms of slow sorption of organic chemicals to natural

particles, Environ. Sci. Technol. 30 (1995) 1-11.

[186] X. Li, X. Cao, G. Wu, T. Temple, F. Coulon, H. Sui, Ozonation of diesel–fuel

contaminated sand and the implications for remediation end-points, Chemosphere 109 (2014)

71-76.

[187] P.A. Hong, S. Nakra, C.J. Kao, D.F. Hayes, Pressure-assisted ozonation of PCB and

PAH contaminated sediments, Chemosphere 72 (2008) 1757-1764.

42

Figure captions:

Fig. 1. Hydroxyl radicals formed according to advanced oxidation technologies.

Fig. 2. The accumulated numbers of scientific papers devoted to the application of advanced

oxidation processes to soils remediation. The data are based on the search results from Web of

Science (May 2015).

Fig. 3. Schematic illustration on decomposition of organic compounds by Fenton processes

Fig. 4. Schematic illustration on decomposition of organic pollutants by TiO2 photocatalysis.

43

Table 1

Overview of work done in the remediation of pesticides contaminated soils by AOPs.

Technology Pesticides studied;

(mg/kg soil)

Experimental conditions Important results Reference

Fenton oxidation Pendimethalin; (100) 10 g soil, 0.05 g FeSO4, 50 mL

H2O2 solution (150-360 g/kg soil),

pH 2-3, 40 h.

99% of the pendimethalin was removed by

Fenton oxidation. Fenton oxidation released

BOD, COD, TOC, and nitrate to the solution.

[82]

Modified Fenton

oxidation

(Fe3+-resin

catalyst)

Pentachlorophenol

(PCP); (1000)

60 g soil, 60 mg PCP, 600 mL

deionized water, 3 g Fe3+-resin

catalyst, 0.1 M H2O2, 80 °C, no

pH adjustment, 100 rpm, 1 h.

94% of PCP was removed in 30–50 min.

Fe3+-resin could be reused for at least six

cycles of PCP oxidation without loss in

efficiency.

[85]

Fenton-like

oxidation

Dichlorodiphenyltrichlor

oethane (DDT) and

dichlorodiphenyldichloro

ethylene (DDE); (320)

20 g soil, 400 mL of 0-0.4 mM

EDTA solution, 1–10 g L-1 of

zero-valent iron (ZVI), no pH

adjustment, 800 rpm, 12 h.

An increase of EDTA and ZVI dosages

improved the removal of the contaminants

significantly. EDTA was simultaneously

degraded.

[86]

Soil washing

with surfactant

followed by

Fenton oxidation

DDT; (1500) and DDE;

(500)

Soil washing: 150 g soil, 1.0 L

TX-100 solutions (8.3 g L−1).

Fenton: 250 mL Wastewater, 12

mM FeSO4, 36 mL of 10 M H2O2,

66% DDT and 80% DDE were removed after

three sequential washings. 99 and 95%

degradation efficiencies were achieved for

DDT and DDE after 6 h Fenton oxidation.

[89]

44

pH 5.7, 6 h.

Soil extraction

followed by

Fenton oxidation

Atrazine; (1000) Soil extraction: soils were washed

with ethanol solution (1.5, 3 and 5

vol. %). Fenton: 100 mL sample,

Fe2+: H2O2=1:10, pH 3, 2 h.

About 95% atrazine was removed by soil

flushing. Only 28.1% atrazine in the solution

was degraded in the presence of ethanol after

2 h Fenton oxidation.

[90]

Photo-assisted

Fenton oxidation

Furadan, alachlor,

atrazine, azinphos-methy,

captan; ( 2×10-4 M)

5×10-5 M Fe (III), 1×10-2 M H2O2,

25.0 °C, pH 2.8, UV irradiation

(20W, 300-400 nm), 2 h.

99% of the total pollutants were removed

after 30 min reaction. Intermediate products

such as formate and oxalate appeared in the

early stages of degradation.

[32]

Photocatalytic

degradation

Diuron; (10,50,100) 50 g soil, TiO2 (0.1, 0.5, 1.0 and

2.0 wt%), solar light (2 mW cm−2,

365 nm), 120 h.

The degradation was limited to the first 4 cm

of soil. Both the catalyst and the diuron

concentration show no influence on the

kinetics of diuron degradation.

[95]

Photocatalytic

degradation

Glyphosate; (40) 50 g Soil, photocatalyst (0.1, 0.25,

0.5%, 1, 5 and 10 wt%), sunlight,

2 h.

0.5% of photocatalyst load has the best

photocatalytic degradation activity. The best

moisture content of soil is 30%~50%.

[96]

Photocatalytic

degradation

DDT; (120) TiO2 (0.1, 0.5, 1.0 and 2.0 wt%),

UV irradiation (300 W, 254, 265,

313, and 365 nm), 25 °C, 24 h.

The humic substances (2 wt%) inhibited the

DDT photodegradation. Alkaline medium

accelerated the photodegradation.

[99]

45

Photocatalytic

oxidation with

composite

photocatalyst

Hexachlorocyclohexane

(HCH); (1234)

Catalyst (10, 30, 50, and 70 wt%),

UV irradiation (300 W, main

wavelengths 254, 265, 313, and

365 nm), 12 h.

Over 90% of HCH was removed.

Photocatalytic activities of the photocatalysts

varied with the content of TiO2 in the order of

10% <70% < 50% < 30%. Photodegradation

rate increases with the soil pH.

[101]

Pulsed corona

discharge plasma

PCP; (200) 5.0 g soil, thickness =3.6 mm,

moisture content =20%, 0-50 kV,

0-100 Hz, pH 7.9, 45 min.

The removal efficiency of PCP reached 92%.

Increasing peak pulse voltage or pulse

frequency resulted in higher PCP degradation

efficiency.

[174]

Pulsed corona

discharge plasma

PCP (200-600) and

p-nitrophenol (PNP;

200-600)

5.0 g soil, thickness =3.6 mm,

moisture content =15%, 18 kV, 50

Hz, pH=5.5, 7.9 and 10.5, 45 min.

Alkaline condition was beneficial for both

PNP and PCP degradation. The existence of

the second pollutant presented inhibitive

effects on pollutants degradation.

[178]

46

Table 2

Overview of work done in the remediation of PAHs contaminated soils by AOPs.

Technology PAHs studied;

(mg/kg soil)

Experimental conditions Important results Reference

Fenton oxidation

catalyzed by

chelated ferrous

ion

light PAHs (1600) and

heavy PAHs (1200 )

30 g soil, 100 mL water, 0.5 M

of chelated Fe2+, Fe2+/H2O2

1:100 or 1:50, no pH

adjustment,12 h.

Less than 100 mg/kg of PAHs was remained in

the soils after the treatment. Too high oxidant

doses can result in a decrease in the oxidation

efficiency.

[113]

Fenton oxidation Phenanthrene (700) and

pyrene ( 615)

5 g soil, 50% (v/v) H2O2 (4, 6

or 8 mL), FeSO4 (0.9, 1.3 and

1.8 mL), pH 3-4, 4h.

94% of phenanthrene was removed, while more

than half of pyrene still remained. The most

important factor was the reaction time,

followed by Fe2+, H2O2 concentration and pH.

[ 114]

Fenton oxidation

coupled with a

chelating agent

(CA)

Phenanthrene (500 )

and fluoranthene

( 500 )

H2O2/soil 0.05, Fe3+/soil 0.025,

Fe3+ solution (100 mL/min),

H2O2 and CA solution (160

mL/min), no pH adjustment,

24h.

79.42% of phenanthrene and 68.08% of

fluoranthene were removed. Phenanthrene

(3-aromatic ring) was more readily degraded

than fluoranthene (4-aromatic ring).

[116]

Fenton oxidation Coal tar contains 12

PAHs; (PAHs: 1000)

5 g soil, 25 ml water, 10 mM

FeSO4, 1.0 wt% H2O2, pH 3,

Both total soil porosity and organic content

affect the susceptibility of PAHs to Fenton

[119]

47

Slurry was shaken at room

temperature, 14 days.

oxidation. PAHs with lower molecular weight

tend to more readily degraded by Fenton

oxidation.

Ethyl lactate

(EL) pre-treated

followed by

Fenton oxidation

Phenanthrene,

anthracene,

fluoranthene and

benzo[a]pyrene; (PAHs:

500)

Pretreatment: 5 g soil, 5 mL EL

solution (EL/water=0.6), 150

rpm, 30°C, 6 h.

Fenton: 5 g soil, 1.0 M of 30

wt% H2O2, 1.0 mL of 1.0 M

Fe2+, no pH adjustment, 8 h.

99.54% of total PAHs were removed. The

accumulation of oxygenated-polycyclic

aromatic hydrocarbonswas observed. EL based

Fenton treatment was most appropriately

adopted for soil with native pH >6.2.

[120]

Photocatalytic

degradation

Pyrene; (40) 5 g soil, TiO2 ( 0, 1, 2, 3, and 4

wt%), UV irradiation (20 W,

253.7 nm), 25 °C , 25 h.

44% of pyrene was removed. The removal

efficiency of pyrene increased along with

increasing the light intensity and the content of

humic acids.

[121]

Photocatalytic

degradation

phenanthrene (40) and

pyrene (40)

5 g soil, TiO2 (0, 1, 2, 3, and 4

wt%), UV irradiation (20W,

253.7 nm), 25 °C, no pH

adjustment, 25 h.

The degradation rate of the phenanthrene and

pyrene on soil surfaces was related to their

absorption spectra in soil. The removal

efficiency of PAHs increased along with

increasing the light intensity and the

concentration of humic acids.

[123]

48

Photocatalytic

degradation

motor oil

contaminated soil;

(PAHs: 40)

Triethylamine and

contaminated soil were stirred

for 10 minutes. TiO2 (1 g/L),

UV irradiation (15 W, 300-400

nm), 24 h.

93-99% removal of these PAHs was achieved

in 24 hours. Once removed from the solid

matrix, the concentrated PAHs can be

photocatalytically degraded efficiently.

[124]

Photocatalytic

degradation

phenanthrene, pyrene

and benzo[a]pyrene;

(PAHs: 40)

5 g soil, 0.5-3 wt% TiO2, UV

irradiation (20 W, 254, 310

and 365 nm), pH 4.2; 6.8 and

9.7, 30 °C, 120 h.

Acidic or alkaline conditions facilitate the

photocatalytic degradation rates of the PAHs

Photocatalytic degradation rates of PAHs

followed the order of 254 nm irradiation > 310

nm irradiation > 365 nm irradiation.

[126]

Photocatalytic

degradation

Pyrene; (40) 5, 10, 20 and 40 mg kg−1 of

humic acid, UV irradiation

(1. �07 mW cm−2, 254 nm),

20-30 °C, pH 6.8, 28h.

The removal rate at 30 °C was greater than

those at 25 and 20 °C. The Photocatalytic

degradation mainly occurred within a soil depth

of 1.0-4.0 mm. A low concentration of humic

acid increased the photocatalytic degradation.

[128]

Ozonation acenaphthene,

henanthrene,

anthracene and

fluoranthene; (PAHs:

10 g soil, pH 6.8, O3: 30–50 L

h−1 and 10–30 ppm, 2–15 min.

A high conversion percentage is obtained in the

first minutes of the process. Fluoranthene

showed the highest removal efficiency.

[72]

49

10)

Ozonation Anthracene; (2000) 40 g soil, moisture content

=20%, O3: 16 and 40 mg/L, 0.5

mL/min, no pH adjustment, 90

min.

Organic matter provokes the additionally ozone

consuming. The majority of by-products

formatted react with O3.

[4]

50

Table 3

Overview of work done in the remediation of PCBs contaminated soils by AOPs.

Technology PCBs studied;

(mg/kg soil)

Experimental conditions Important results Reference

Fenton oxidation Aroclor 1242; (100) 1 g soil, 4.6 mL of 5 wt% H2O2,

0.4 mL of the Fe2(SO4)3 solution

(100 ppm), pH 2.75, 25°C, 72h.

98% of the original PCBs were removed after

72 hours treatment. The degree of degradation

was dependent on the level of congener

chlorination.

[134]

Cyclodextrins

modified Fenton

oxidation

2,2’,6,6’-tetrachlorobip

henyl (100) and

3,3’,5,5’-tetrachlorobip

henyl (100)

50 g soil, 1 mM Fe2+, H2O2 (5 mM

h-1), 0.3 mM cyclodextrins, pH 3,

room temperature, 12 h.

Addition of cyclodextrins increased the

degradation efficiency of PCBs. Cyclodextrins

chelated the iron, allowing the Fenton reaction

to be carried out at near neutral pH.

[136]

Fenton and

photo-Fenton

oxidation

Aroclor 1242; (100) Fenton: 1 g soil, 4.6 mL H2O2 (1,

5 and 10%), Fe3+ (100, 500 and

1500 ppm), 15, 30 and 50 °C, pH

2.75, 72 h. Photo-Fenton: UV

irradiation (20W, 254 nm), 4h.

Over 85% of the total PCBs were removed

after 72 hours Fenton oxidation. Close to 100%

of the total PCBs were removed by

Photo-Fenton oxidation in 30 minutes.

[26]

Modified Fenton

oxidation

soils contaminated with

PCBs were collected

10 g soil, 10 mL H2O2 (2-50

wt% ), 10 mM of the stabilizers

98% of the total PCBs were removed. Using

high H2O2 concentrations is appropriate for the

[138]

51

from polluted sites;

(PCBs: 120)

phytate, citrate, or malonate, 10

mM iron (III)-EDTA, no pH

adjustment, 24 h.

treatment.

Photocatalytic

degradation

Contaminated

sediment; (PCBs:

218-228)

1.5 L slurry (80 g soil), 1.5 g

TiO2, UV irradiation (100 W, 365

nm), no pH adjustment, 92 h.

Up to 81% of the total PCBs were removed.

Removal efficiency decreased when highly

chlorinated aroclors or terphenyls are present.

[141]

photocatalytic

degradation with

added surfactant

3,3,4,4-tetrachlorobiphe

nyl; (100)

0.4 g soil, 100 mL water, 500

mg/L TiO2, 1% surfactant, UV

irradiation (21W, 285-315 nm),

48 °C, no pH adjustment, 48 h.

PCB degradation rates in samples followed the

order spiked clay > spiked soil > River bank

soil. PCB was difficult to release from

aged-contaminated soil.

[143]

soil washing

followed by

photocatalytic

degradation

2,4,4’-trichlorobiphenyl

; (100)

Soil washing: 0.5 g soil, 10 mL of

extracting solution, 150 rpm, 2 h.

Photocatalysis: 40 mL oil washing

solution, 500 mg/L TiO2, UV

irradiation (525µW cm−2, 365

nm), no pH adjustment, 8 h.

The extracting percentage was significantly

affected by the chlorination degree of PCBs.

Polyoxyethylene lauryl ether was suitable for

treating PCB-contaminated soil since it

supported the feasibility of both soil washing

and photocatalytic degradation.

[144]

soil washing

followed by

photocatalytic

soils contaminated with

PCBs were collected

from a waste

500 mL soil washing solution, 200

mg TiO2, UV irradiation (100 W,

300 W, 254, 292, 313, 334, 365,

The degradation rates were 71% and 87% for

systems with TiO2 and graphene-TiO2,

respectively. PCBs molecules were

[145]

52

transformer factory 436 and 546 nm), pH 1-13, 20 °C,

2 h.

dechlorinated gradually to biphenyl and then

decomposed to small molecule.

Ozonation PCB congeners; (not

given)

40 g soil, O3: 60 g m−3, 0.45 L

min−1, no pH adjustment, 6 h.

Ozonation was more efficient for PCBs

degradation in freshly spiked soils. Soil pH

decreased after the treatment.

[182]

Pressure-assisted

ozonation

Waukegan Harbor

sediment; (PCBs: 5.1)

Soil slurry (100–1000 mL,

0.1–0.4 w/w), O3: 1.5% by

volume, 1 L min−1, ultrasonic

irradiation (600 W, 20 kHz), pH

7.4–7.9, 0.5 h.

PCBs were completely removed. The

confluence of O3 and PCBs at the interface has

thus resulted in the accelerated removal of the

contaminants.

[187]

53

Table 4

Overview of work done in the remediation of TPHs contaminated soils by AOPs.

Technology TPHs studied;

(mg/kg soil)

Experimental conditions Important results Reference

Fenton oxidation Soil was sampled from

an oil sludge

contaminated site;

(TPHs: 11198)

3 kg soil, 13 wt% H2O2, 10 mM

Fe2+, 6 L distilled water, 20 h at

pH 6.5, 20 h at pH 4.5, and 40 h at

pH 3.0.

Fenton oxidation was efficient in degrading

the oil contaminants in the soil. The two steps

stabilization processes were necessary to

enhance environmental protection and to

render final product economically profitable.

[155]

Fenton-like

oxidation

Diesel; (TPHs: 1000) 5 g soil, 6 kinds of iron catalysts

(5 to 25 mM), 5 mL of H2O2 (0.15

to 1.5 M), no pH adjustment, 90 h.

Over 99% of diesel was degraded by iron (III)

perchlorate and iron (III) nitrate catalyzed

Fenton reaction. Iron (III) sulfate, iron (II)

sulfate and iron (II) perchlorate provided

70–80% diesel oxidation

[157]

Basic oxygen

furnace slag (BOF

slag) catalyzed

Fenton-like

oxidation

No. 6 fuel oil and

diesel; (TPHs: 10000)

50 g soil, 30 mL H2O2 (0-30

wt%), BOF slag (0, 100, 200, 300,

400, and 500 g kg−1), no pH

adjustment, 40 h.

76% and 96% of fuel oil and diesel were

removed, respectively, at the optimal

conditions (15% of H2O2 and 100 g kg−1 of

BOF slag). The oxidation of TPHs was

enhanced with the addition of BOF slag.

[159]

54

Electrokinetic-

Fenton oxidation

Diesel; (TPHs:10000) 13 kg soil, 5500 mL of electrolyte

(tap water, 0.01 M NaCl, or 0.1 M

NaCl solution) and H2O2 solution

(4 and 8 wt% of H2O2), no pH

adjustment, 45 d.

Electrokinetic-Fenton oxidation obtained

higher TPHs removal efficiency (97%)

compared to the efficiencies observed from

electrokinetic oxidation (55%) or Fenton

oxidation (27%) alone.

[163]

Soil washing

followed by

electro-Fenton

oxidation

Contaminated soils

from urban site;

(TPHs:6100)

Soil washing: 15 kg soil, 1-5%

Tween 80 (3 mL min-1).

electro-Fenton: 0.15 mM Na2SO4,

1000 mA, 20 °C, pH 3, 32 h.

The efficiency of the soil washing treatment

was very low (only 1% after 24 h of

washing). Over 99.5% of TPHs in the eluates

was removed after the electro-Fenton

oxidation.

[164]

Fenton-like

oxidation coupled

with a chelating

agent

Diesel; (TPHs:10000) 5 g soil, 10 ml water, 20 mmol L−1

Fe3+, 4000 mmol L−1 H2O2, 5

mmol L−1 sodium citrate, 20 °C,

no pH adjustment, 7 h.

The oxidant is stabilized by sodium citrate,

which allows the treatment applied at natural

pH (7.22). About 37 % of TPHs was removed

after the treatment.

[165]

Fenton

pre-treatment

followed by

biodegradation

Weathered petroleum

oil-contaminated soil;

(TPHs: 38300)

Fenton: 2 kg soil, 30 wt% H2O2,

H2O2/Fe(III)-NTA 50:1, pH 7.5,

24 h.

Biodegradation: 2 kg soil, 200 g

peanut hull, C:N:P ratio of

After bioremediation for 20 weeks, reduction

of TPHs by 88.9% was observed in the

combined treatment compared with 55.1% in

the biological treatment alone. The activity of

microbial communities also increased by

[167]

55

100:10:5, 20 weeks. Fenton pre-treatment.

Fenton-like

pre-treatment

followed by

biodegradation

transformer oil and

shale oil; (TPHs:

20000)

Fenton-like: 15 g soil, 15 mL

H2O2 (30 wt%), 20 °C, pH 3.0 or

6.7, 72 h.

Biodegradation:.No microbial

inoculums were added, 20 °C, no

pH adjustment, 30 d.

The acidic pH (3.0) conditions favoured

Fenton-like oxidation; nevertheless,

remediation of contaminated soil using in situ

Fenton-like treatment was more feasible at

natural soil pH. Combined chemical and

biological processes were more effective than

either one alone.

[168]

Graded modified

Fenton’s (MF)

oxidation

pre-treatment

followed by

biodegradation

tank oil; (TPHs: 4840) Fenton: 10 g soil, 40 mL water, 10

mL iron catalyst (6.98 mM Fe2+),

2.5 mL of 30 wt% H2O2 (5 times),

natural pH, 25 h.

Biodegradation: 10 g soil, 50 mL

phosphate buffer, 6mL

macro-elements, and 0.6 mL

micro-element, 26 d.

Three-time addition of H2O2 was found to be

favorable and economical due to decreasing

TPHS removal from three time addition

(51%) to five time addition (59%). Removal

efficiency of tank oil was up to 93% after four

weeks, with a 31% increase comparing to

non-oxidized soil.

[170]

Photocatalytic

degradation

Spent motor oil (55.6

mL/ kg)

90 g soil, 2 g TiO2, 5 ml oil,

sunlight (640 W/m2), 40 h.

The highest degree of oil decomposition was

observed during the first hours. Photocatalytic

degradation occurs only on the soil surface.

[171]

56

Dielectric barrier

discharge plasma

Kerosene; (7400) 2 g soil, thickness =2 mm,

moisture content =15%, 15-20 kV,

40 Hz, no pH adjustment, 22 min.

The total kerosene components abatement can

reaches 90%. The main removal mechanism

is the oxidation of kerosene in the soil matrix.

[172]

Ozonation Diesel; (25000) 15 g soil, moisture content =20%,

O3: 10, 30 and 50 mg/L, 180

mL/min, pH 6.0, 20 h.

Soil moisture below 18% did not influence

the ozonation efficiency. Soil pH declined

from 6 to 3.

[186]

Figure 1

Figure 2

Figure 3

Figure 4

57

Graphical abstract:

58

Highlights

•The main drawback of conventional Fenton treatment is the reduction in soil pH.

•Modified Fenton treatments can produce •OH at a pH near neutral.

•The lack of visible light activity hinders the practical applications of photocatalysis.

•Ozonation is suitable to treat soils with large pore spaces and low moisture.

•Plasma oxidation is able to treat soils with high concentration pollutants.

1) We have received the above-mentioned article for publication in which equations

are provided in plain text mode. Kindly provide the equations in Math type/equation

editor format.

Response: The equations in Math type/equation editor format are provided as follow:

OHHOFeOHFe 322

2•++→+

−++ (1)

111 sM 70pH3, −−

=K

oxidationFurther ROHOHRH 2 →•+→•+ (2)

oxidationFurther ROHOHR →•→•+ (3)

•++↔++++

22

223 HOHFeOHFe (4)

114 sM 0.1-0.001 −−

=K

( ) OHOHFeOHFe 222

•+→++

+2

(5)

( ) OHFeOHFe 22•+→+

++

hv (6)

OH2OH 22 •→+ hv (7)

( )−+

++•→+ eHOHMOHM 2 (8)

+−

+→+ heTiO 2 hv (9)

( ) OHHTiOOHhTiO 222 •++→+++ (10)

( ) OHTiOHOhTiO 22 •+→+−+ (11)

−−

•→+ 22 OeO (12)

•→+•+−

22 HOHO (13)

( ) 22222 TiOOHeTiOHHO +→++•−+ (14)

( ) 2222 TiOHOOHeTiOOH ++•→+−− (15)

•+→+++ RHhRH (16)

2322 OOHHOOOH +•+→+− (17)

OHHOeOH 22 •+→+−− (18)

23 OHO-SoilSoilO +•→+ (19)

2) In the supplied manuscript there is citation for table[5] in your paper, but we have

not received the table[5]. Please e-mail the table[5] to me so that we may continue

with the publication of your paper.

Response: We are very sorry for the mistakes. Table 1 in the original manuscript has

been removed according to one of the reviewers’ comments. Table 2-5 in appendix of

the original manuscript has been revised to Table 1-4. Unfortunately, we forgot to

change the serial number of the Tables in the text.

- Line 219 “Table 2” should be “Table 1”;

- Line 327 “Table 3” should be “Table 2”;

- Line 443 “Table 4” should be “Table 3”;

- Line 525 “Table 5” should be “Table 4”.