hydroxyl radicals based advanced oxidation processes (aops) for remediation of soils contaminated...
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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
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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).
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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.
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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
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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
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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
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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)
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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
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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
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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
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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
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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
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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
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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
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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).
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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”.