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Advanced Oxidation Processes in Water/Wastewater Treatment:

Principles and Applications. A Review

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DOI: 10.1080/10643389.2013.829765

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Critical Reviews in Environmental Science and Technology, 44:2577–2641, 2014
Copyright © Taylor & Francis Group, LLC
ISSN: 1064-3389 print / 1547-6537 online
DOI: 10.1080/10643389.2013.829765

Advanced Oxidation Processes

in Water/Wastewater Treatment: Principles
and Applications. A Review

MEHMET A. OTURAN and JEAN-JACQUES AARON

Laboratoire Géomatériaux et Environnement, Université Paris-Est, UPEM,
Marne-la-Vallée, France

Advanced oxidation processes (AOPs) constitute important, promis-

ing, efficient, and environmental-friendly methods developed to
principally remove persistent organic pollutants (POPs) from wa-
ters and wastewaters. Generally, AOPs are based on the in situ
generation of a powerful oxidizing agent, such as hydroxyl rad-
icals ( •OH), obtained at a sufficient concentration to effectively
decontaminate waters. This critical review presents a precise and
overall description of the recent literature (period 1990–2012) con-
cerning the main types of AOPs, based on chemical, photochemical,
sonochemical, and electrochemical reactions. The principles, per-
formances, advantages, drawbacks, and applications of these AOPs
to the degradation and destruction of POPs in aquatic media and
to the treatment of waters and waste waters have been reported and
compared.

KEY WORDS: anodic oxidation, AOPs, COT, degradation prod-

ucts, electro-Fenton, Fenton’s reagent, hydroxyl radicals, mineral-
ization, photocatalysis, photo-Fenton, toxicity assessment, water
treatment

Address correspondence to Mehmet A. Oturan, Laboratoire Géomatériaux et Envi-

ronnement, Université Paris-Est, UPEM, EA 4508, 77454 Marne-la-Vallée, France. E-mail:
mehmet.oturan@univ-mlv.fr
Color versions of one or more of the figures in the article can be found online at
www.tandfonline.com/best.

2577
2578 M. A. Oturan and J.-J. Aaron

1. INTRODUCTION

Water is essential to life, and, although it is very abundant on Earth, it

is mainly constituted by aquatic resources which are not directly usable
by human beings, such as salted waters of oceans and seas (97.2% of the
water total mass), and glaciers (2.15%). Only about 0.65% of the water to-
tal mass can be directly utilized by human. Moreover, the distribution of
water is geographically very unequal, and some regions are quasi desert,
and an important part of water resources are more or less polluted. There-
fore, there is a crucial need of developing efficient and ecologically friendly
methods to treat contaminated waters and reduce or completely eliminate
pollutants.
Among these methods, advanced oxidation processes (AOPs) have been
precisely defined by Glaze et al. (1987) as water treatment processes per-
formed at room temperature and normal pressure and based on the in situ
generation of a powerful oxidizing agent, such as hydroxyl radicals (•OH),
at a sufficient concentration to effectively decontaminate waters. AOPs have
been recently the object of increasing attention, as shown by the large num-
ber of fundamental and applied research works (Andreozzi et al., 1999;
Herrmann et al., 1999; Tarr, 2003; Gogate and Pandit, 2004; Parsons, 2004;
Brillas et al., 2006; Laine and Cheng, 2007; Zaviska et al., 2009). Indeed,
they constitute promising, efficient and environmental-friendly methods to
remove persistent organic pollutants (POPs) from waters. Several types of
AOPs are based on the in situ formation of •OH radicals by means of var-
ious chemical, photochemical, sonochemical, or electrochemical reactions.
The oldest and most used chemical AOP is the Fenton method, in which a
mixture of a soluble iron(II) salt and H2 O2 , known as the Fenton’s reagent,
is applied to degrade and destroy POPs (Andreozzi et al., 1999; Tarr, 2003;
Gogate and Pandit, 2004; Parsons, 2004). However, it is possible to greatly
improve the oxidation efficiency and utilization easiness of this method by
simultaneously irradiating the treated environmental water sample by ul-
traviolet (UV) light (photo-Fenton method) or sunlight (solar photo-Fenton
method) (Tarr, 2003). Also, other photochemical methods, such as heteroge-
neous photocatalysis, using TiO2 suspensions (Herrmann et al., 1999; Kon-
stantinou and Albanis, 2003; Cernigoj et al., 2007), as well as ozonolysis
(O3 + UV irradiation) (Rosenfeldt et al., 2006), have been reported. An-
other interesting improvement is based on combining the Fenton technique
with electrochemical reactions. In fact, there exist a number of electrochem-
ical advanced oxidation processes (EAOPs) which have been recently re-
ported. (Oturan, 2000; Brillas et al., 2006, 2009; Martı́nez-Huitle and Ferro,
2006; Panizza and Cerisola, 2009b). EAOPs have been applied to destroy
persistent or toxic organic matter in water by in situ generation of •OH
via electrochemistry. EAOPs constitute emergent and environmental-friendly
techniques since they use a clean reagent, the electron, avoiding or reducing
 Advanced Oxidation Processes in Water/Wastewater Treatment 2579

considerably the use of chemical reagents. •OH is generated either directly

by oxidation of water on a high O2 evolution overvoltage anode (anodic ox-
idation (AO)) (Panizza and Cerisola, 2009), or indirectly in bulk solution by
using electrochemically generated Fenton’s reagent from electrode reactions
(Brillas et al., 2009).
To the best of our knowledge, relatively few reviews (Oturan et al.,
2004; Parsons, 2004; Brillas et al., 2006, 2008; Martı́nez-Huitle and Ferro,
2006; Pignatello et al., 2006; Brillas and Oturan, 2007; Oturan and Brillas,
2007; Zaviska et al., 2009; Wang and Xu, 2012) have been devoted until now
to the principles and development of AOPs and their applications for water
treatments. The goal of this review is to present a precise and overall descrip-
tion of the recent literature (period 1990–2012) concerning various types of
AOPs, based on chemical, photochemical, sonochemical, and electrochem-
ical reactions, their performances and their application to the degradation
and destruction of toxic and/or POPs in aquatic media and the treatment of
waters and waste waters.

2. CHEMICAL AOPs
2.1 Fenton’s Reagent
The Fenton’s chemistry started as early as the end of the nineteenth century,
when Fenton published, in a pioneering work, a detailed study on the use
of a mixture of H2 O2 and Fe2+ (later called the Fenton’s reagent) for the
oxidation and destruction of tartaric acid (Fenton, 1894).
Because of its important development during the twentieth century and
a number of applications to water and soil treatment, several review papers
have been focused on Fenton’s chemistry (Merli et al., 2003; Neyens and
Baeyens, 2003; Ikehata and El-Din, 2006; Pignatello et al., 2006; Bautista
et al., 2008).
In the 1930s, Haber and Weiss (1932, 1934) have shown that the cat-
alytic decomposition of H2 O2 by iron salts obeyed to a complex radical and
chain mechanism. More recent mechanistic studies have demonstrated that
the Fenton process was initiated by the formation of hydroxyl radical, in
agreement with the classical Fenton’s reaction (1), and could be applied to
the degradation/destruction of various organic pollutants (Metelitsa, 1971;
Sun and Pignatello, 1993; Gallard et al., 1998):

Fe2+ + H2 O2 → Fe3+ +• OH + OH− (1)

Since the reaction (1) takes place in acidic medium, it can be also written
as:

Fe2+ + H2 O + H+ → Fe3+ + H2 O +• OH (2)

2580 M. A. Oturan and J.-J. Aaron

The formation of •OH radical in the Fenton process, supported by Walling

(1998), has been experimentally confirmed by means of chemical probes
and/or spectroscopic techniques (Lindsey and Tarr, 2000). Several rate con-
stants involved in the Fenton’s chemistry have been also measured by means
of pulse radiolysis (Rush and Bielski, 1985).
The Fenton process can be efficiently applied when the pH optimum
value of the polluted aqueous medium is about 2.8–3.0. Indeed, in these
conditions, the Fenton’s reaction can be propagated by the catalytic behavior
of the Fe3+/Fe2+ couple. It is worthwhile to note that only a small catalytic
amount of Fe2+ is required, since this ion is regenerated from the so-called
Fenton-like reaction (3) between Fe3+ and H2 O2 : (Haber and Weiss, 1932,
1934):

Fe3+ +H2 O2 → Fe2+ +HO•2 +H + (3)

Compared to •OH, formed HO2 • radical is characterized by a lower oxi-

dization power and, therefore, is significantly less reactive towards organic
compounds (Bielski et al., 1985). It has been reported that the Fenton-like
reaction (3) is much slower than the Fenton’s reaction itself (reaction 1)
(Brillas et al., 2009). In fact, Fe2+ can be more rapidly regenerated by the
reduction of Fe3+ with HO2 • from reaction (4) (Merli et al., 2003), with an
organic radical R• from reaction (5) and/or with superoxide ion (O2 •−) from
reaction (6) (Rothschild and Allen, 1958).

Fe3+ + HO•2 → Fe2+ + O2 + H+ k2 = 2 × 103 M−1 s−1 (4)

Fe3+ + R→ Fe2+ + R+ (5)
Fe 3+
+ O•−
2 → Fe 2+
+ O2 k2 = 5 × 10 M s
7 −1 −1
(6)

In spite of the complexity of its mechanism, the Fenton process was applied
to oxidation and degradation/destruction of organic pollutants as soon as
the mid-1960s (Brown et al., 1964), and many applications rapidly devel-
oped, essentially in the 1990s and the 2000s (Barbeni et al., 1987; Kuo, 1992;
Potter and Roth, 1993; Li et al., 1997; Tang and Tassos, 1997; Watts et al.,
1997; Gallard et al., 1998; Wang et al., 1999; Yoshida et al., 2000; Gogate
and Pandit, 2004; Pignatello et al., 2006). Indeed, a number of studies have
demonstrated the efficiency of the Fenton process in many cases. For exam-
ple, the Fenton’s reagent has been used for treating wastewater (Gogate and
Pandit, 2004), for discoloring effluents of dye industries (Kuo, 1992), and
for destroying toxic organic compounds, such as 2,4,6-trinitrotoluene (TNT)
(Li et al., 1997), 2,4-dinitrophenol (Wang et al., 1999), chlorophenols (Bar-
beni et al., 1987; Potter and Roth, 1993), chlorobenzenes (Watts et al., 1997),
tetrachloroethylene (Yoshida et al., 2000), and haloalkanes (Tang and Tassos,
1997).
 Advanced Oxidation Processes in Water/Wastewater Treatment 2581

Also, it is important to stress that the practical efficiency of the Fenton’s

reagent strongly depends on various factors such as temperature, pH, H2 O2 ,
and catalyst concentrations, which control the regeneration capacity of Fe2+
from Fe3+ produced during the process and the rate of oxidation of organics
by generated •OH.
Basically, the Fenton process possesses several important advantages
for water/wastewater treatment (Bautista et al., 2008):

(i) A simple and flexible operation permitting easy implementation in ex-

isting plants;
(ii) Easy-to-handle and relatively inexpensive chemicals;
(iii) No need for energy input.

Nevertheless, the following drawbacks have been also noted (Tarr,

2003):

(i) Rather high cost and risks due to the storage and transportation of H2 O2 ;
(ii) Need of important amounts of chemicals for acidifying effluents at pH
2–4 before decontamination and/or for neutralizing treated solutions
before disposal;
(iii) Accumulation of iron sludge that must be removed at the end of the
treatment;
(iv) Impossibility of overall mineralization due to the formation of Fe(III)-
carboxylic acid complexes, which cannot be efficiently destroyed with
bulk •OH.

Some remedies have been proposed to minimize these disadvantages. For

instance, it is possible to reduce the added amount of H2 O2 by optimiz-
ing its concentration. Moreover, iron sludge may be prevented by the use
of solid iron-containing catalysts, including zeolites, alumina, iron-modified
clays, mesoporous molecular sieves, iron oxides, ion-exchange resins, or
iron-exchanged Nafion membranes, easily separated from the treated solu-
tions (Pignatello et al., 2006; Bautista et al., 2008). For applications to the
treatment of actual wastewaters, the Fenton method has been also coupled to
various processes, such as coagulation, membrane filtration, and biological
oxidation, allowing to degrade more extensively organic pollutants (Mantza-
vinos and Psillakis, 2004; Lucas et al., 2007). Finally, another possibility is
the application of the photo-Fenton process, which will be discussed in
Section 3.

2.2 Peroxonation
The principle of peroxonation is based on a coupling between ozone (O3 )
and H2 O2 , resulting in the generation of oxidizing radicals. As pointed out by
2582 M. A. Oturan and J.-J. Aaron

Zaviska et al. (2009), the peroxonation process should be more efficient than
ozonation alone, since H2 O2 has the effect of increasing the decomposition
rate of O3 in water, which produces a larger number of very reactive •OH
radicals.
The mechanism and conditions of application of peroxonation have
been investigated by Paillard et al. (1988), who have shown that a very fast
reaction occurred between H2 O2 under its ionized form (HO2 −, pK a = 11.6)
and ozone, leading to the formation of •OH radicals:

O3 + HO− •
2 → O2 + OH + O2
−•
(7)

HO2 • radicals are also obtained by the reaction of •OH radicals with HO2 −.
Then, all these radicals can decompose H2 O2 by other mechanisms occurring
under optimum experimental conditions (pH = 7.7 and H2 O2 /O3 ratio =
0.5) (Paillard et al., 1988).
The peroxonation process has been successfully applied by several au-
thors (Chromostat et al., 1993; Paillard, 1994) to the elimination of microp-
ollutants and toxic compounds (hydrocarbons, pesticides . . . ) found in in-
dustrial waters, drinkable waters, and ground waters. The oxidation system
by O3 /H2 O2 can be inserted between a filtration on sand and a filtration on
active coal in a reactor through which water is running. The main goals of
the water treatment by peroxonation are to significantly lower the microp-
ollutant concentration before filtration on active coal, in order to increase
the active coal filter lifetime. It is worthwhile to note that the H2 O2 /O3 ratio
should be kept constant in all points of the reactor and that the H2 O2 residual
concentration should not be above the maximum value of 0.5 mg L−1 in the
treated waters.
The practical usefulness of the peroxonation process is limited by sev-
eral factors, such as the low water solubility of ozone, the important energetic
consumption, and its sensitivity to several factors, including the pH, tempera-
ture, micropollutant type, and the occurrence of side reactions which are also
consuming •OH radicals like the other AOPs (Buxton et al., 1988; Hernan-
dez et al., 2002). Nevertheless, the principal advantages of the peroxonation
system are that it is simple to handle and it has a great bactericide activity.
For these reasons, this method has been developed as an essential step of
disinfection for the treatment of drinkable waters. For instance, Galey and
Paslawski (1993) have applied peroxonation to eliminate several pesticides,
including phenylureas, organochlorines (lindane and endosulfan), and tri-
azines (atrazine, simazine, and terbutryne), from wastewater plants. Initially,
the wastewaters contained a concentration of about 0.1 μg L−1 for each pes-
ticide, and, after treatment by the peroxonation process, between 80% and
90% of the pesticides were destroyed.
 Advanced Oxidation Processes in Water/Wastewater Treatment 2583

3. PHOTOCHEMICAL AOPs

The photochemical technologies present the advantages to be simple, clean,

relatively inexpensive, and generally more efficient than chemical AOPs.
Also, they can disinfect waters, and destroy pollutants. Consequently, UV
radiations have been coupled with powerful oxidants such as O3 and H2 O2 ,
including, in some cases, a catalysis with Fe3+ or TiO2 , resulting in various
kinds of important photochemical AOPs. These photochemical processes
are able to degrade and/or destroy pollutants by means of three possible
reactions, including photodecomposition, based on UV irradiation, excita-
tion and degradation of pollutant molecules, oxidation by direct action of
O3 and H2 O2 , and oxidation by photocatalysis (with Fe3+ or TiO2 ), inducing
the formation of •OH radicals. In this section, we have examined the prin-
ciples and applications of these different AOPs, namely the H2 O2 photolysis
(H2 O2 /UV), O3 photolysis (O3 /UV), photo-Fenton process (H2 O2 /Fe2+/UV),
and heterogeneous photocatalysis (TiO2 /UV). We have presented in Table 1
a comparison of the experimental conditions, applications, and performances
of these various photochemical AOPs for some selected examples.

3.1 Photolysis of H2 O2 (H2 O2 /UV)

Hydrogen peroxide can be photolyzed by UV radiations absorbed at wave-
lengths ranging from 200 to 300 nm, yielding the homolytic scission of the
O–O bond of the H2 O2 molecule and leading to the formation of •OH radi-
cals which can also contribute to the decomposition of H2 O2 by secondary
reactions (Hernandez et al., 2002; Zaviska et al., 2009). In these conditions,
several successive and competitive reaction steps can take place:

H2 O2 + hv → 2• OH (8)
•
OH + H2 O2 → H2 O + HO•2 (9)
HO2 • + H2 O2 →• OH + H2 O + O2 (10)
•
OH + HO− •
2 → HO2 + OH
−
(11)
•
2HO2 → H2 O2 + O2 (12)
•
OH + HO•2 → H2 O + O2 (13)
2• OH → H2 O2 (14)

Equation (8) corresponds to the initiation step, Eqs. (9)–(11) to the propaga-
tion steps, and Eqs. (12)–(14) to the termination steps.
It is worthwhile to stress that the rate of production of free radicals
mainly depends on different important parameters, including the charac-
teristics of UV lamps (emission spectrum, power . . . ) and physicochemical
 TABLE 1. Comparison of the results and performances of different photochemical advanced oxidation processes (AOPs) for selected examples

2584
Type of AOP Type of water Pollutant Experimental (optimal) conditions Performances/remarks References
H2 O2 /UV Simulated reactive Monochlorotriazine type UVC; [H2 O2 ] = 680 mg L–1; pH = TOCa removal = 30.4%; EE/Ob = Alaton et al. (2002)
dyebath effluents reactive dyes 3.0 0.633 kWh m–3
Distilled water; Six azo dyes: Acid UVC; [H2 O2 ] = 240 mg L–1; pH = Color removal = 95% at time = Shu and Chang (2005)
wastewater Orange 10, Acid Red 5.3; [Azo dye]0 = 20.0 mg L–1 26–92 min; Consumed energy =
14 and 18; Acid 2141 × 10−3–7666 × 10−3 kWh,
Yellow 17; Direct according to the dye
Yellow 4; Acid Black
1
Deionized water Five chlorophenoxy acid UVC; [H2 O2 ] = 170 mg L–1; pH = Total photodegradation time = Fdil et al. (2003)
herbicides: MCPA, 7.0 20–90 min; Mineralization yield
MCPP, 2,4-D, 2,4-DP, (from CODd) = 56–79%,
2,4,5-Tc according to the herbicide
O3 /UV Distilled water; Six azo dyes: Acid UVC; O3 flow rate = 6.0 dm3 min–1; Color removal = 95% at Shu and Chang (2005)
–1
wastewater Orange 10, Acid Red [Azo dye]0 = 20.0 mg L time < 11.5 min; Consumed
14 and 18; Acid energy =
Yellow 17; Direct 349 × 10−3–954 × 10−3 kWh,
Yellow 4; Acid Black according to the dye
1
Deionized water Two endocrine UVC; O3 flow rate = Complete photoconversion time = Irmak et al. (2005)
disrupters: 17β 7.6 × 10−3–15.9 × 10−3 mmol min–1, 45 min for E2 (0.715 mmol O3 ) and
–estradiol (E2 ) and according to the compound 75 min for BPA (1.4 mmol O3 )
bisphenol A (BPA)
Distilled water; Three pesticides: UVC; O3 flow rate = 1.2 g/h; pH = Pesticide removal = 92–96% (t = Lafi and Al-Qodah (2006)
wastewater Deltamethrin, 7.0; T = 25 ◦ C; C0 = 100 mg L–1 210 min); CODd removal =
Lambda-cyhalothrin, 90–95% (with biological treatment),
Triadimenol according to the compound
2+
Photo- Simulated textile R94H reactive UVC; [H2 O2 ] = 100 mg/L; [Fe ] = Color removal = 96% at time = Kang et al. (2000)
–1
Fenton wastewater dye + PVA 20 mg L ; pH = 3–5; [Dye]0 = 30 min; CODd removal = 36%
100 mg L–1 (60 min)
Deionized water Azo-dye Orange II UVA; [H2 O2 ] = 200 mg/L; [Total Discoloration ∼ 100% at irradiation Maezono et al. (2000)
Fe] = 8.0 mg L–1; pH = 3.0; time = 15 min
–1
[Dye]0 = 60 mg L
−3
Deionized ultra-pure Azo dye mixture: Acid UVA; [H2 O2 ] = 8 × 10 M; Complete discoloration of the dye Macı́as-Sánchez et al. (2011)
water Yellow 36 and Methyl [Fe2+] = 3 × 10−4 M; pH = 2.0; mixture sample at time = 75 min;
–1
Orange [Dye]0 = 50 mg L (in dye TOCa removal (mineralization) =
mixture) 100% in 180 min—Incomplete
decomposition and TOCa removal
for the Fenton process
 Deionized water Sulfamethazine (SMT) UVA; [H2 O2 ] = 600 mg L–1; [Fe2+] = Total SMT removal in 2 min; TOCa Pérez-Moya et al. (2010)
antibiotic 40 mg L–1; pH = 3; T = removal: 50% in less than 30 min
18–19 ◦ C; [SMT]0 = 50 mg L–1
Deionized water Five chlorophenoxy acid UVC; [H2 O2 ] = 40 mM; [Fe3+] = Total photodegradation time = Fdil et al. (2003)
herbicides: MCPA, 4 mM; pH = 3.0 7–60 min; Mineralization yield
MCPP, 2,4-D, 2,4-DP, (from CODd) = 80–96%,
2,4,5-Tc according to the herbicide—Much
shorter degradation times and
larger mineralization yields than
for the H2 O2 /UV AOP
Distilled water; natural Abamectin pesticide UVA; [H2 O2 ] = 6 mM; [Fe3+] = Pesticide removal = 80% for Kaichouh et al. (2008)
water 0.5 mM; pH = 2.5; C0 = distilled water and 70% for natural
9.0 mg L–1 water (t = 60 min); TOCa
removal = 60% in 180 min—Only
40% of pesticide removal (t =
60 min) for the Fenton process
Deionized water; tap 3-Chloropyridine (ClPy)e UVA (UV and solar light); [H2 O2 ] = TOCa removal (mineralization) = Ortega-Liébana et al. (2012)
water 8.8 mM; [Fe2+] = 0.88 mM; pH = 100% in 60 min (UV lamp) and
2.8; T = 25 ◦ C; [ClPy]0 = 40 ppm 120 min (solar light)—Only 22% of
mineralization (t = 120 min) for
the Fenton process
TiO2 /UV Simulated reactive Monochlorotriazine type UVA; [TiO2 ] = 103 mg L–1; pH = 7.0 TOCa removal/1 hr treatment = Alaton et al. (2002)
dyebath effluents reactive dyes 10.3%—Discoloration =
94.6%—No electrical cost (solar
energy)
3
Deionized water Imazalil (Imaz) fungicide UVA; [TiO2 ] = 2.5 × 10 mg/L; Complete removal of Imaz within Hazime et al. (2012)
pH = 6.5–10; [Imaz]0 = 35 min of irradiation; TOCa
25 mg L–1 removal (mineralization) = 100%
in ∼ 800 min
e
Deionized water 3-Chloropyridine (ClPy) UVA (UV and solar light); [TiO2 ] = TOCa removal (mineralization) = Ortega-Liébana et al. (2012)
–1
700 mg L ; pH = 6.8; [ClPy]0 = 100% in 300 min (UV light)—TiO2
40 ppm photocatalysis process about
5 times slower than photo-Fenton
process
Deionized water Pharmaceutical agent UVA (UV and solar simulator); Total abatement of salbutamol within Sakkas et al. (2007)
–1
Salbutamol [TiO2 ] = 649 mg L ; pH = 6.8. 30 min of irradiation; complete
mineralization within 180 min of
irradiation
aTOC = total organic carbon.
bEE/O = Electrical energy per order of pollutant removal.
cMCPA = 4-chloro-2-methylphenoxyacetic acid; MCPP = 2-(2-methyl-4-chlorophenoxy)propionic acid; 2,4-D = 2,4-dichlorophenoxyacetic acid; 2,4-DP =

2-(2,4-dichlorophenoxy)propionic acid; 2,4,5-T = 2,4,5-trichlorophenoxyacetic acid.

dCOD = chemical oxygen demand.

2585
e3-Chloropyridine chosen as a model compound of pyridine pesticides.
2586 M. A. Oturan and J.-J. Aaron

properties of the medium (pH, transmission of UV radiations, turbidity . . . )

(Crissot, 1996). Generally, the reaction rate is larger in alkaline medium at
pH > 10, which can be attributed to the fact that the HO2 − anion (Eq. (13)),
resulting from the ionization of H2 O2 , can strongly absorb UV radiations and
produce free radicals (HO2 • and •OH). However, a drawback of this AOP is
that the molar absorption coefficient of H2 O2 is relatively weak in the UV
region, and, consequently, it is necessary to use a rather strong concentration
of hydrogen peroxide for an efficient oxidation of organic pollutants.
The H2 O2 /UV AOP has been applied to decontaminate ground wa-
ters (Eckenfelder et al., 1992), and to eliminate cyanides and organic pollu-
tants, like benzene, trichloroethylene, tetrachloroethylene, etc. (Doré, 1989).
In a review, Ikehata and El Din (2006) have investigated the efficiency of
H2 O2 /UV and photo-Fenton-type AOPs for the degradation in aqueous me-
dia of eight main groups of pesticides, including aniline derivatives, car-
bamates, chlorophenoxy acids, organochlorinated derivatives, organophos-
phates, pyridine and pyrimidine derivatives, triazines, and substituted ureas.
Alaton et al. (2002) have compared the treatment efficiency of the H2 O2 /UV
process to that of other AOPs (O3 /OH−, and TiO2 /UV-A) for the oxi-
dation of simulated reactive dye bath effluents containing a mixture of
monochlorotriazine-type dyes and various dye auxiliary chemicals at typical
concentrations encountered in exhausted reactive dye bath liquors (Table 1).
Also, Colonna et al. (1999) have studied the decolorization and mineraliza-
tion of some azo- and anthraquinone dyes by UV irradiation in the presence
of hydrogen peroxide.

3.2 Photolysis of O3 (O3 /UV)

Ozone in aqueous solution absorbs UV radiations between 200 and 360 nm,
with a maximum at 253.7 nm (molar absorption coefficient ε max =
3600 L mol−1 cm−1) (Van Craeynest et al., 2004). Since the ε max value of O3 is
much larger than that of H2 O2 at this wavelength (ε max = 18.6 L mol−1 cm−1),
the ozone photolysis process should be more efficient in these conditions
(low-pressure Hg lamp—λ = 253.7 nm) than H2 O2 photolysis. Therefore,
the O3 /UV process has been, compared with other oxidation processes,
widely applied to the environment, particularly in the treatment of waters
and waste waters for eliminating toxic POPs, such as pesticides and phe-
nolic compounds (Masten et al., 1997; Trapido et al., 1997; Trapido and
Kallas, 2000; Aaron and Oturan, 2001; Rosenfeldt et al., 2006; Cernigoj et al.,
2007).
The photolysis of ozone in water leads to the formation of •OH radicals,
which are very reactive and efficient oxidizing species, according to the fol-
lowing successive and competitive steps (Van Craeynest et al., 2004; Zaviska
 Advanced Oxidation Processes in Water/Wastewater Treatment 2587

et al., 2009):

O3 + H2 O + hv → 2 • OH + O2 (15)
•
O3 + OH → HO•2 + O2 (16)
O3 + HO•2 → OH + 2 O 2
•
(17)
•
OH + HO•2 → H2 O + O2 (18)
2 • OH → H2 O2 (19)

In this reaction scheme, Eq. (15) corresponds to the initiation step, Eqs. (16)
and (17) to the propagation steps, and Eqs. (18) and (19) to the termination
steps.
This AOP based on the ozone photolysis has been particularly utilized
to eliminate various volatile chlorinated organic compounds (VCOCs). For
instance, the usefulness of the O3 /UV AOP for the oxidation of several
VCOC, including CHCl3 , CCl4 , trichloroethylene (TCE), tetrachloroethylene,
and 1,1,2-trichloroethane (TCA), has been investigated by Bhowmick and
Semmens (1994). It was found that direct ozonation allowed to oxidize
CHCl3 , while the action of hydroxyl radicals oxidized CHCl3 , TCA, and TCE,
and that CCl4 could not be destroyed by direct ozonation or by the •OH
radicals (Bhowmick and Semmens, 1994). TCE was also decomposed by the
O3 /UV AOP, using an hybrid pilot reactor equipped with an air pollution
control system, whereas there was no effect on nonchlorinated VOC (Striebig
et al., 1996). Effluents containing various types of organic pollutants, such
as pesticides (Aaron and Oturan, 2001; Lafi and Al-Qodah, 2006), endocrine
disrupters (Irmak et al., 2005; Lau et al., 2007), pharmaceutical compounds
(Ikehata et al., 2006; Gebhardt and Schroeder, 2007), antibiotics (Akmehmet
and Otker, 2004), surfactants (Amat et al., 2007), dyes (Shu and Chang, 2005;
Yonar et al., 2005; Wu and Chang, 2006; Hsing et al., 2007), and nitroben-
zene (Tong et al., 2005), were decontaminated (Table 1). A good example
of application of the O3 /UV AOP is the work of Irmak et al. (2005) on the
degradation of two endocrine disrupters, namely 17β-estradiol and bisphe-
nol A (BPA) in aqueous medium by using ozone and ozone/UV techniques.
Figure 1 presents the experimental setup used for decomposition and com-
plete degradation of both endocrine disrupters. In Figure 2, we have shown
the effect of different O3 dosages on the conversion rates of BPA under UV
irradiation. As can be seen, the complete BPA (0.10 mmol) oxidation by
the O3 /UV process was achieved by using about 18.7 × 10−3 mmol min–1
O3 dosage within 75 min, which corresponds to 1.4 mmol of O3 . In all
dosages, the BPA oxidation rates were found to be faster in the O3 /UV pro-
cess than with simple ozonation, which shows the superiority of the O3 /UV
AOP.
2588 M. A. Oturan and J.-J. Aaron

FIGURE 1. Experimental setup for the ozonization (O3 /UV) technique, used for decompo-
sition and complete degradation of two endocrine disrupters, namely 17 β–estradiol and
bisphenol A. (Source: Irmak et al., 2005)
C Elsevier. Reproduced by permission of Elsevier. Permission to reuse must be obtained from
the rightsholder.

3.3 Photo-Fenton (H2 O2 /Fe2+/UV)

It was shown, in a pioneering work of Zepp et al. (1992), that the classical
Fenton reaction (1) could be photo-assisted by using UV radiations to stim-
ulate the catalytic reduction, in H2 O2 aqueous solutions, of Fe3+ into Fe2+,

FIGURE 2. Decrease of bisphenol A (BPA) concentration (in mmol) with time during appli-
cation of the O3 /UV process at different O3 dosages. (Source: Irmak et al., 2005)
C Elsevier. Reproduced by permission of Elsevier. Permission to reuse must be obtained from
the rightsholder.
 Advanced Oxidation Processes in Water/Wastewater Treatment 2589

which increases the formation of •OH radicals, according to reaction (20):

Fe3+ + H2 O + hu → Fe2+ + H+ +• OH (20)

At pH 2.8–3.5, the preeminent form of Fe3+ is the [Fe(OH)]2+ ion which plays
a key role in this so-called photo-Fenton process (Pignatello et al., 2006).
The formation of •OH radicals by photo-Fenton reactions has been quantified
in aqueous solutions containing Fe(III)-oxalate complexes and H2 O2 (Zepp
et al., 1992). Moreover, in the photo-Fenton process, UV irradiation has also
the ability to directly decompose H2 O2 molecules into hydroxyl radicals, like
in the H2 O2 /UV process (Eq. (8)).
In fact, the photo-Fenton process can use several UV regions as light
energy source, namely UVA (λ = 315–400 nm), UVB (λ = 285–315 nm),
and UVC (λ < 285 nm). It is worthwhile to note that the intensity and wave-
length of UV radiations has a significant effect on the destruction rate of
organic pollutants. However, a drawback of this process is the important
economical cost arising from the utilization of artificial light. But, an alterna-
tive approach, recently developed, consists to use sunlight (at wavelengths
λ > 300 nm) as free and renewable energy source in the so-called solar
photo-Fenton process, based on solar collectors, for photocatalytic decon-
tamination and/or disinfection of waters (Malato et al., 2007; Oller et al.,
2007; Silva et al, 2007). Indeed, this solar photo-Fenton process seems to be
a much more satisfactory method than classical lamp-driven photo-Fenton
both from the economic and environmental standpoints, as indicated by the
application of the recently proposed, easy-to-use environmental-economic
index.
As already pointed out, the action of photons in photo-Fenton pro-
cess is quite complex. In fact, the classical Fenton’s reaction (1), in which
hydroxyl radicals are produced, presents the inconvenience of a large ac-
cumulation of Fe3+ species, decelerating the efficiency of treatment. This
drawback is avoided in the photo-Fenton process, since the reductive pho-
tolysis of [Fe(OH)]2+, (reaction (20)), has the advantage to regenerate the
Fe2+ ions that catalyze Fenton’s reaction (1) and to yield additional •OH
radicals (Faust and Hoigné, 1990; Pignatello, 1992):
Quantum efficiency values of 0.04 ± 0.04 at 313 nm and 0.017 ± 0.003
at 360 nm (293 K, ionic strength = 0.03 M) were estimated for reaction (21)
(Faust and Hoigné, 1990). More recently, a novel kinetic method, based on
the use of dimethylsulfoxide (DMSO) as a •OH probe compound, has been
developed for the determination of the quantum yields for the photolysis of
Fe(III)-hydroxy complexes, including [Fe(OH)]2+ (Lee and Yoon, 2004). The
individual quantum yield values for the photolysis of the monomeric Fe(III)
complexes were found to decrease with increasing wavelength in the range
240–380 nm (Lee and Yoon, 2004).
2590 M. A. Oturan and J.-J. Aaron

In addition, UV light can induce the photodegradation of some oxida-

tion by-products or their complexes with Fe(III) promoting Fe2+ regener-
ation (Zuo and Hoigné, 1992; Safarzadeh-Amiri et al., 1996). For example,
the kinetics and mechanism of decomposition of oxalic acid by sunlight-
induced photolysis in atmospheric water of Fe(III)-oxalate complexes, such
as Fe(C2 O4 )+, Fe(C2 O4 )2 −, and Fe(C2 O4 )3 3−, that can absorb photons from
250 to 580 nm, have been investigated in detail by Zuo and Hoigné (1992).
The authors found that the photolysis (photodecarboxylation) of the Fe(III)-
oxalate complexes in de-aerated solution followed the overall reaction (21):

2Fe(C2 O4 )(3−2n)+
n + hv → 2Fe2+ + (2n − 1)C2 O2−
4 + 2CO2 (21)

It has been postulated that, in the corresponding photochemical mechanism,

the absorption of a photon by an Fe(III)-oxalate complex leads to an excited
state in which it occurs a ligand-metal electron transfer (Zuo and Hoigné,
1992).
The classical photo-Fenton process has been successfully applied in
a number of studies to the degradation and/or mineralization of various
pollutants such as pesticides (Pignatello, 1992; Sun and Pignatello, 1993;
Fallmann et al., 1999; Fdil et al., 2003; Kaichouh et al., 2004; Mestankova
et al., 2004; Kaichouh et al., 2008; Diagne et al., 2009; Boufia-Chergui et al.,
2010), antibiotics (Pérez-Moya et al., 2010; Rozas et al., 2010), dyes (Kang
et al., 1999, 2000; Xie et al., 2000; Maezono et al., 2010; Macı́as-Sánchez et al.,
2011), chlorophenols and other chlorinated compounds (Pupo Nogueira and
Guimarães, 2000; Ghaly et al., 2001), and TNT residues (Liou et al., 2004), in
waters. This photo-Fenton process has been also widely used for the treat-
ment and remediation of contaminated wastewaters and landfill leachates
(Primo et al., 2008; Hermosilla et al., 2009; Kondo et al., 2010) (Table 1).
In several studies, the efficiency of photo-Fenton AOP was compared
with that of other photochemical oxidation processes (Ghaly et al., 2001;
Fdil et al., 2003; Primo et al., 2008; Macı́as-Sánchez et al., 2011). For in-
stance, Fdil et al. (2003) have investigated the efficiencies of three oxida-
tion processes, including UV irradiation alone, H2 O2 /UV, and photo-Fenton
(H2 O2 /Fe2+/UV) processes, for the degradation and mineralization of five
chlorophenoxyalcanoic herbicides, and they have noted a similar behavior
for all compounds. Indeed, when applying UV irradiation alone, the pho-
todegradation of the chlorophenoxyalcanoic herbicides under study was very
slow and partial (1.5–3.0 hrs, according to the compound) and did not allow
to completely destroy the photoproducts, with small mineralization yields of
35–69%, whereas, in the case of the other two AOPs, the photodecomposi-
tion was much faster (30 min–more than 60 min for H2 O2 /UV, and 6–60 min
for photo-Fenton, according to the compound), and the mineralization yields
were much higher (56–79% for H2 O2 /UV, and 80–96% for photo-Fenton, ac-
cording to the compound). This indicated that the photo-Fenton process
was by far the best method for the rapidity of photodecomposition as well
 Advanced Oxidation Processes in Water/Wastewater Treatment 2591

FIGURE 3. Mineralization of 2,4,5-trichlorophenoxyacetic acid (2,4,5-T) herbicide during ap-

plication of UV irradiation alone, and of the H2 O2 /UV and photo-Fenton (H2 O2 /Fe2+/UV)
processes in terms of (a) COD and (b) released Cl− ions. (Source: Fdil et al., 2003)

C Groupement d’Intérêt Scientifique des Sciences de l’Eau (GIS Eau), France. Reproduced
by permission of Groupement d’Intérêt Scientifique des Sciences de l’Eau (GIS Eau), France.
Permission to reuse must be obtained from the rightsholder.

as for the mineralization yield of all herbicides. As an example, Figure 3 com-

pares the mineralization curves of 2,4,5-trichlorophenoxyacetic acid (2,4,5-T)
by the three oxidation processes. Also, Ghaly et al. (2001) have compared
the photochemical oxidation of p-chlorophenol by the UV/H2 O2 and photo-
Fenton processes. The authors found out that the photo-Fenton process was
the most effective treatment process under acidic conditions and yielded a
higher rate of degradation of p-chlorophenol in a very short radiation time. It
accelerated the oxidation rate by 5–9 times relative to the UV/H2 O2 process.
In another study, Primo et al. (2008) have investigated the performances of
several oxidation processes for the organic matter removal in the treatment
of landfill leachates, and they have observed that the removal efficiencies de-
creased in the order: photo-Fenton > Fenton-like > Fenton > H2 O2 /UV > UV
alone. Therefore, they have recommended the use of photo-Fenton process
as the most efficient AOP. Macı́as-Sánchez et al. (2011) have compared the
performances of the photo-Fenton process and the chemical Fenton reaction
for the degradation of a model azo dye mixture. They found that, with the
photo-Fenton AOP (λ = 365 nm), decolorization of the dye mixture sample
was achieved within 70 min and complete mineralization, evaluated by TOC
abatement, was completed in 180 min, whereas, using the Fenton reaction,
only 75% of mineralization was obtained in the same time. These results
again demonstrate that the photo-Fenton process was much more efficient
than the Fenton one for the decolorization and mineralization of an azo dye
mixture.
Moreover, the solar photo-Fenton AOP, which constitutes, as previously
indicated, an interesting alternative to the classical photo-Fenton process,
has been recently utilized for the elimination of various organic compounds
2592 M. A. Oturan and J.-J. Aaron

present in natural or polluted waters (Ortega-Liébana et al., 2012), degra-

dation of herbicides (Silva et al., 2007, 2010), and treatment of wastewater
effluents and landfill leachates (Klamerth et al., 2010; Vilar et al., 2011). For
example, 15 emerging organic contaminants at low concentrations were de-
graded in simulated and real effluents of municipal wastewater treatment
plant, using a solar photo-Fenton process at fixed pH and with a Fe con-
centration of 5 mg L−1 in a pilot-scale solar reactor (Klamerth et al., 2010).
In another applied study, Vilar et al. (2011) developed a solar photo-Fenton
process combined with a biological nitrification and denitrification system
for the decontamination of a landfill leachate in a pilot plant based on pho-
tocatalytic and biological devices (immobilized biomass reactor). Also, solar
collectors to degrade water contaminants by photocatalysis and to inacti-
vate microorganisms present in waters have been reviewed by Malato et al.
(2007), and a pilot plant with a coupled solar photocatalytic photo-Fenton-
biological system to enhance the biodegradability and complete mineral-
ization of a bio-recalcitrant industrial compound, α-methylphenylglycine, in
distilled water and simulated seawater (500 mg L−1), has been successfully
built (Oller et al., 2007).

3.4 Heterogeneous Photocatalysis (TiO2 /UV)

In the early seventies, Fujishima and Honda (1972) showed the possibility
of using the photo-excited semiconductor titanium dioxide (TiO2 ) to split
water into hydrogen and oxygen in a photo-electrochemical solar cell. As
pointed out in several reviews (Mills and Le Hunte, 1997; Fujishima et al.,
2000; Pelaez et al., 2012), their fundamental work led to the development of
a new AOP technology, based on semiconductor photocatalysis, for numer-
ous environmental and energy applications. This so-called heterogeneous
photocatalysis involves irradiation with near-UV light a semiconducting cat-
alyst, generally TiO2 , preferably in its rutile in front anatase form, which is
easily photo-excited to form electron-donating and electron-accepting sites,
permitting to induce oxidation–reduction reactions. When the absorbed UV
photons have an energy larger than the energy gap (between the valence
and the conducting bands) of the semiconductor, electron-hole pairs are
formed, which can either recombine or migrate to the semiconductor sur-
face and then react with chemical species adsorbed on the surface (Zaviska
et al., 2009).
Titanium dioxide has been mostly chosen for the application of hetero-
geneous photocatalysis processes to water treatment because it is a material
close to being a practically ideal photocalyst in several important aspects.
First, TiO2 is highly stable chemically and biologically inert, very easy to
produce, inexpensive, active from the photocatalysis standpoint, and it has
an energy gap comparable to that of solar photons (Pignatello, 1992; Zaviska
 Advanced Oxidation Processes in Water/Wastewater Treatment 2593

FIGURE 4. Schematic band diagram showing the potentials for several redox processes
occurring on the TiO2 surface at pH 7. (Source: Fujishima et al., 2000)
C Elsevier. Reproduced by permission of Elsevier. Permission to reuse must be obtained from
the rightsholder.

et al., 2009). Moreover, the photogenerated holes are strong oxidants, and
the photogenerated electrons are reducing enough to yield superoxide from
dioxygen. The energy band diagram for TiO2 is presented in Figure 4. As
can be seen, the redox potential for photogenerated holes is 2.53 V versus
the standard electrode hydrogen (SHE). In these potential conditions, the
photogenerated holes are able to either directly oxidize the absorbed pollu-
tants or oxidize the hydroxyl groups located at the TiO2 surface to form •OH
radicals, whose redox potential is only slightly decreased (Fujishima et al.,
2000). Consequently, the degradation of pollutants contained in the contam-
inated waters can take place either directly at the semiconductor surface or
indirectly through interactions with the •OH radicals, the indirect oxidation
by the radicals being the most favored degradation pathway. In addition,
it is possible to again increase the number of •OH radicals by adding into
the photoreactor H2 O2 or O3 which can be photolyzed by UV irradiation
(Zaviska et al., 2009).
During the heterogeneous photocatalytic process, the TiO2 catalyst can
be utilized either under dispersed form (powder, aqueous suspension) or in
thin film form (fixed TiO2 catalytic layer). The Fujishima group (Fujishima
et al., 2000) has widely participated to the preparation of TiO2 films, by
putting TiO2 coatings on various types of support materials. The dispersed
TiO2 catalyst presents several advantages: it is easy to use, it possesses an
important specific surface, and it can be aerated which prevents the recom-
bining of electron-hole pairs and increases the catalyst efficiency. However,
a drawback of the dispersed form is the progressive formation of dark cat-
alytic sludge, which diminishes the efficiency of UV irradiation and reduces
the photoreactor performances. In contrast, for TiO2 films, there is no need
2594 M. A. Oturan and J.-J. Aaron

to separate the catalytic particles at the end of the process, but the catalytic
layer must be very stable and active. Also, the amount and type of catalyst
to be used depend on the irradiation source, the nature and concentration
of pollutant to be treated, and the photoreactor. Moreover the pH value of
the medium plays a crucial role in the efficiency of photocatalysis and must
be optimized in a preliminary step, according to the type of pollutant under
treatment. For example, in the case of weak acid pollutants, the photocatal-
ysis efficiency increases when the pH diminishes, yielding a decrease of the
polarity of the pollutant which is more easily adsorbed at the catalyst surface
(Zaviska et al., 2009).
The heterogeneous TiO2 photocatalysis has been widely applied in re-
cent years, particularly in the case of organic pollutants refractory to oxida-
tion by the other conventional AOPs (Mills and Le Hunte, 1997; Zaviska et al.,
2009). Also, it is able to completely destroy pathogenic biologic pollutants,
including viruses, bacteria, and mold (Zaviska et al., 2009). This technology
is generally very efficient for treating a substantial range of inorganic as well
as organic pollutants. Moreover, it is worthwhile to point out the recent de-
velopment of various strategies to modify TiO2 for the use of visible light
(visible light active TiO2 photocatalytic materials), including nonmetal and/or
metal doping, dye sensitization, and coupling semiconductors (Pelaez et al.,
2012).
A large number of applications of heterogeneous TiO2 photocatalysis,
particularly in the field of water purification, have been recently described
(Mills and Le Hunte, 1997; Fujishima et al., 2000; Pelaez et al., 2012). For
example, toxic, inorganic ions, such as cyanide, bromate, nitrite, and sulfite,
have been oxidized by this process into nontoxic or weakly toxic com-
pounds (CO2 , bromide, nitrate, sulfate) (Mills and Le Hunte, 1997; Zaviska
et al., 2009). Heterogeneous TiO2 photocatalytic degradation and/or min-
eralization were performed in the case of a number of organic pollutants,
such as pesticides (Herrmann et al., 1999; Konstantinou and Albanis, 2003;
Cernigoj et al., 2007; Hazime et al., 2012; Rivera-Utrilla et al., 2012; Seck et al.,
2012), pharmaceuticals (Sakkas et al., 2007), surfactants such as dodecylben-
zenesulfonate (Sanchez et al., 2011), sulfur-containing organic compounds,
dyes (Lin et al., 2012), and chloropyridine (Ortega-Liébana et al., 2012).

4. SONOCHEMICAL AOPs

Ultrasounds in aqueous medium constitute a particular AOP technology

which can proceed through two distinct types of actions, either a chemi-
cal (indirect) or a physical (direct) mechanism. In the indirect action, gen-
erally realized at high frequency, water and dioxygen molecules undergo
homolytic fragmentation and yield •OH, HO2 •, and •O radicals (Riez et al.,
1985; Lorimer and Mason, 1987; Trabelsi et al., 1996; Zaviska et al., 2009).
 Advanced Oxidation Processes in Water/Wastewater Treatment 2595

The direct action, so-called sonication, involves the formation by the ultra-
sounds of cavitation bubbles which grow, then collapse, creating powerful
breaking forces with extremely high temperatures (T = 2000–5000 K) and
pressures (about 6 × 104 kPa). In these extreme conditions, a sonolysis of
water molecules occurs, which produces very reactive radicals able to react
with organic chemical species present in the aqueous medium (Eqs. (22) and
(23)), and/or a pyrolysis degradation of organic compounds is taking place
(Eq. (24)) (Hua and Hoffmann, 1997; Ma, 2012):

H2 O + ))) → • OH +• H (22)
•
OH + X (O.C.) → Products (23)
X (O.C.) + H → Products (24)

where: ))) = ultrasound, X(O.C.) = organic compound(s).

Therefore, in recent years, ultrasounds have been widely used for the oxi-
dation and degradation/destruction of organic pollutants in waters, wastew-
aters, and sewage sludges, as shown in several studies (Naffrechoux et al.,
2003; Minero et al., 2005; Dai et al., 2006; Ku et al., 2006; Kim and Huang,
2007; Torres et al., 2007; Sun et al., 2007; Qiu et al., 2008; Ghodbane and
Hamdaoui, 2009; Zaviska et al., 2009; Ma, 2012). However, a notable draw-
back of the water treatment by ultrasounds is that the number of generated
•
OH radicals is generally insufficient. Consequently, ultrasounds have been
more often applied in the presence of other oxidants, such as H2 O2 and
dioxygen, as well as in combination with UV irradiation and with various
AOPs, including the Fenton’s reagent (with different forms of iron: Fe0, Fe2+,
and Fe3+) and Fenton-type reactions. The later combined methods are gen-
erally called sono-Fenton AOPs. In these conditions, the development of
such advanced hybrid techniques has resulted in the improvement of the
degradation/destruction efficiency of organic pollutants in waters and of the
reduction of the sonochemical treatment time (Trabelsi et al., 1996; Naffre-
choux et al., 2003; Minero et al., 2005; Dai et al., 2006; Liang et al., 2007; Sun
et al., 2007; Torres et al., 2007; Namkung et al., 2008; Oturan et al., 2008b;
Qiu et al., 2008; Ghodbane and Hamdaoui, 2009; Joseph et al., 2009; Ma
et al., 2010; Ma and Sung, 2010; Ma, 2012), although a clear understanding
of the theory and mechanisms of these various sonochemical hybrid AOP
techniques needs to be done.
As pointed out and described in detail in the very recent review of Ma
(2012), several important, operational parameters, including the ultrasonic
frequency (range: 20–1700 kHz) and amplitude, the iron type (Fe0, Fe2+,
and Fe3+) and dosage, and the solution pH, must be optimized in order
to improve the efficiency of the sonochemical AOP treatment of polluted
waters. Moreover, it has been also shown that the optimal values of ultrasonic
frequency and amplitude depended on the characteristics of the effluents
2596 M. A. Oturan and J.-J. Aaron

to be treated (nature and concentration of pollutants . . . ) as well as on

the experimental conditions (type and volume of the sonochemical reactor,
treatment time . . . ), and that the physicochemical properties of effluents
(vapor pressure, surface tension, viscosity, presence of impurities and/or
gas . . . ) had a significant effect on the performances of the sonochemical
reactor (Gogate and Pandit, 2000, 2001, 2004; Zaviska et al., 2009).
In fact, during the last decade, the sonochemical AOPs have been
broadly applied to the degradation/destruction of a great variety of pol-
lutants, including pesticides (Oturan et al., 2008a; Ma et al., 2010; Ma and
Sung, 2010), dyes (Minero et al., 2005; Sun et al., 2007; Ghodbane and Ham-
daoui, 2009; Ma, 2012), aromatic compounds (Trabelsi et al., 1996; Petrier
et al., 1999; Dai et al., 2006; Ku et al., 2006; Liang et al., 2007; Namkung
et al., 2008; Ma, 2012), endocrine disrupters (bisphenol A) and pharmaceu-
ticals (Torres et al., 2007; Ma, 2012), and disinfectant by-products (Kim and
Huang, 2007) in waters and/or wastewaters.
Although the combination of ultrasounds with the Fenton-type reactions
has resulted into the rapid and recent development of sonochemical methods
for the removal of organic pollutants from waters, and seems able to lead
to a very promising technologic approach for decontamination purposes,
most experimental works have been performed until now at the laboratory
scale in artificial systems, with only one or two compounds as the model
contaminants (Ma, 2012). Therefore, application of sonochemical AOPs at the
industrial level in a real-time water (or wastewater) treatment plant would
be needed to demonstrate the economic and commercial feasibility of these
methods.

5. ELECTROCHEMICAL AOPs

Electrochemistry is a technique based on the transfer of electrons, which

makes it particularly interesting from the environmental point of view, since
it constitutes a clean and effective way to produce in situ hydroxyl radical
(•OH) which are able to destroy a large variety of toxic and POPs. These
•
OH radicals can be electrochemically produced either directly (via AO pro-
cess) or indirectly via in situ electrocatalytically generated Fenton’s reagent
(electro-Fenton (EF) process). The effectiveness of the process can be further
increased by combining both electrochemical processes, namely the simul-
taneous AO with boron-doped diamond (BDD) anode and the classical EF
process (carbon felt or gas diffusion cathodes) (Oturan, 2000; Brillas et al.,
2009).
These EAOPs using direct electrochemistry (AO) or indirect electro-
chemistry (EF) oxidations have the needed requirements for decontamina-
tion/detoxification water purposes, because these processes:
 Advanced Oxidation Processes in Water/Wastewater Treatment 2597

— allow rapid degradation of organic pollutants while preventing the

formation of new toxic species;
— lead to total mineralization of organic pollutants;

— use few or no chemical reagents;

— have energy costs as low as possible.

The EAOPs have been mainly developed during the last decade, and have
received great attention, due to their environmental safety and compatibility
(operating at mild conditions), versatility, high efficiency, and possibility of
automation. One of the major advantages of electrochemistry is its ability
to control and produce in situ hydroxyl radicals without adding chemical
reagents or large amounts of catalyst in the medium, allowing the treated
effluents to be directly discharged in natural waters (Oturan, 2000; Brillas
et al., 2009; Nidheesh and Gandhimathi, 2012; Sires and Brillas, 2012).
In this section, the principles, essential features, and recent develop-
ments of the two main EAOPs, namely AO and EF processes, and their
couplings with other photochemical (photoelectro-Fenton (PEF) and so-
lar PEF), sonochemical (sonoelectro-Fenton), and physicochemical (peroxi-
coagulation) treatment methods, will be reviewed. We will also illustrate the
large capacity of oxidation and mineralization of these EAOPs for the treat-
ment and destruction of organic (industrial, agricultural, and pharmaceutical)
pollutants with various examples.

5.1. Anodic Oxidation

As already stated, AO constitutes a direct and clean way to electrochemically
generate •OH radicals, without using any chemicals, the unique reagent be-
ing the electrons. •OH radicals are directly formed at the anode surface by
oxidation of water, with high O2 evolution overvoltage anodes (Comninellis
and De Battisti, 1996: Simond et al., 1997; Perret et al., 1999; Panizza et al.,
2001; Canizares et al., 2006a, 2006b; Brillas and Martinez-Huitle, 2011). Ini-
tially, the AO process was developed with Pt, PbO2 , doped SnO2 , IrO2 , or
DSA (dimensionally stable anodes) that are mainly mixed metal oxide anodes
(Tahar and Savall, 1998, 1999; Boye et al., 2003b; Wu et al., 2012). Marselli
et al. (2003) have proposed a catalytic mechanism, including the generation
of heterogeneous hydroxyl radicals M(•OH) by water electrolysis (reaction
(25)) and oxidation of organics (reaction (26)), based on their experimental
results:

M + H2 O → M(• OH) + H+ + e− (25)

•
M( OH) + R → M + mCO2 + nH2 O + pX (26)
2598 M. A. Oturan and J.-J. Aaron

where M = anode material; M(•OH) = heterogeneous •OH radicals

adsorbed on the anode material; R = organic matter and X = inorganic
ions.
Unlike the metallic anodes, which do not lead to the formation of sig-
nificant amount of oxidants, the metal oxides and DSA anodes were found
to facilitate the formation of chlorine from aqueous chloride ions, which
enhanced their oxidation efficiency because of the formation of hypochlor-
ous acid (HClO), a relatively strong oxidant. The positive effect of mediated
oxidation via HClO has been demonstrated, for example, in the removal of
atrazine (Malpass et al., 2006) dye basic Blue 3 (Özcan et al., 2008a), thio-
carbamate herbicides (Mogyoródy, 2006), methamidophos (Martı́nez-Huitle
et al., 2008), isothiazolin (Han et al., 2011), etc. Due to the natural abundance
of chloride in polluted waters, chlorine-mediated oxidation is widely em-
ployed for in situ generation of oxidants, such as ozone, persulfate (S2 O8 2−),
and HClO with respective standard oxidation potential of 2.07, 2.01, and
1.67 V/SHE, for wastewater treatment according to the following anode re-
actions (Panizza and Cerisola, 2009):

2Cl− → Cl2 + 2e− (27)

+ −
Cl2 + H2 O → HClO + H + Cl (28)
3H2 O → O3 + 6H+ + 6e− (29)
−
4 → S2 O8 + 2e
2SO2− 2−
(30)

The oxidants, obtained by AO of the corresponding anions (reactions

(27)–(30)), were used to oxidize the pollutants present in solution. More-
over, the metal oxide or mixed metal oxide anodes generally exhibit high
oxidation rates but poor mineralization yields. Indeed, the oxidation power
of chlorine and other anode-formed oxidants species remains low compared
to hydroxyl radicals, and, consequently, do not allow to efficiently transform
several intermediates into CO2 and H2 O.
The performances of metallic anodes, particularly the Pt anode, have
been assessed in the removal of chlorophenoxy herbicides (Brillas et al.,
2004), paracetamol (Brillas et al., 2005), and chloroxylenol (Skoumal et al.,
2008), and compared to those of a new and powerful anode, namely the
BDD thin film anode. Metallic anodes have been shown to possess a weaker
oxidation and/or mineralization power than BDD anode for the treatment of
the organic pollutants under study.
Indeed, this BDD anode, which recently appeared as an emerging and
very promising anode material, has very good properties for the electro-
chemical treatment of wastewaters contaminated by organic pollutants. It
possesses a great chemical and electrochemical stability, a wide electrochem-
ical working range and a great oxidation/mineralization power compared to
other anodes. Moreover, this new anode has an O2 overvoltage much higher
 Advanced Oxidation Processes in Water/Wastewater Treatment 2599

than that of conventional anodes such as Pt, PbO2 , doped SnO2 , and IrO2
(Comninellis and De Battisti, 1996; Boye et al., 2002b; Canizares et al., 2004).
Due to this high O2 overvoltage, the BDD anode is able to yield, by reac-
tion (26), larger amounts of •OH radicals physisorbed on the anode surface,
namely BDD (•OH), that are more reactive than the •OH radicals produced by
other anode materials. Thus BDD anode is able to generate large quantities
of reactive heterogeneous hydroxyl radicals from water and other oxidants,
such as O3 , S2 O8 2−, HClO, P2 O8 2, C2 O6 2−, etc., from various ions typically
present in water (SO4 2−, Cl−, PO4 3−, CO3 2−, etc.), and consequently to un-
dergo the direct and mediated oxidation of organic pollutants (Panizza and
Cerisola, 2003b; Cañizares et al., 2009; Rodrigo et al., 2010).
Regarding the direct oxidation, the BDD (•OH) produced at the an-
ode surface leads to a rapid and efficient destruction of organic pollutants.
However, since the generated hydroxyl radicals are adsorbed on the anode
surface, the oxidation process is mass transfer controlled. Therefore, in the
case of organic pollutant low concentration, the process efficiency is not
very high whereas with mediated mechanisms (presence of sulfate, chlo-
rine, phosphate . . . salts), the oxidation reactions can simultaneously occur
through direct (electrode surface) and mediated (bulk) processes which sig-
nificantly increase the global efficiency of this AOP.
The great effectiveness of BDD (•OH) action for the oxidation of a wide
range of organic pollutants has been demonstrated in several studies (Ini-
esta et al., 2001; Panizza and Cerisola, 2003b; Canizares et al., 2008; Özcan
et al., 2008b; Rodrigo et al., 2010), almost yielding the complete mineraliza-
tion of treated solutions. It was highlighted in these studies that the current
efficiency was strongly influenced by the applied current density and initial
pollutant concentration, higher mineralization yields being favored by low
organic pollutant concentrations and high applied current densities. Con-
sequently, the electrochemical oxidation with BDD anode of wastewaters
containing large variety of organic pollutants has been carried out. The AO
was successfully applied to the assessment of polluted waters containing
various pesticides, such as chlorophenoxy herbicides (Brillas et al., 2004),
amitrole (Da Pozzo et al., 2005), parathion (Pedrosa et al., 2006), meco-
prop (Sirés et al., 2008), propham (Özcan et al., 2008b), methamidophos
(Martı́nez-Huitle et al., 2008), chlorpyrifos (Samet et al., 2010), atrazine (Bor-
ras et al., 2010), and pesticide mixtures (Kesraoui-Abdessalem et al., 2010a,
2010b), as well as pharmaceutical residues (Brillas et al., 2005; Guinea et al.,
2008, 2010; Isarain-Chavez et al., 2011; Bensalah et al., 2012; Dominguez
et al., 2012; El-Ghenymy et al., 2012), phenol and chlorophenols (Rodrigo
et al., 2001; Canizares et al., 2003), synthetic dyes (Panizza and Cerisola,
2007, 2008; Hammami et al., 2008), polyaromatic compounds (Panizza and
Cerisola, 2003a), surfactants (Panizza and Cerisola, 2003a; Panizza et al., 2005;
Louhichi et al., 2008; Saez et al., 2010), landfill leachates (Panizza et al., 2010),
car wash wastewaters (Panizza and Cerisola, 2010a, 2010b), and tannery
2600 M. A. Oturan and J.-J. Aaron

effluents (Panizza and Cerisola, 2004). In these studies, the different operat-
ing parameters were investigated, and in all cases almost complete removal
of the pollutants under study was reached with high mineralization effi-
ciencies. Table 2 gives more details on a selected number of examples of
applications carried out by AO.
In addition, the electrochemical process efficiency can be improved
by coupling it either to light irradiation (Skoumal et al., 2008; Malpass et al.,
2012) or to ultrasounds (Garbellini et al., 2010). In the first case, the formation
of supplementary hydroxyl radicals was improved, while in the second one,
the mass transfer rate towards anode was enhanced.

5.2 EF Process
Over the last decade, EAOPs based on cathodic electrogeneration of hydro-
gen peroxide and catalytic regeneration of Fe2+ were developed and suc-
cessfully applied for the treatment of wastewaters containing several families
of persistent and or toxic organic pollutants (Oturan, 2000; Oturan et al.,
2004, 2009b; Oturan and Brillas, 2007; Brillas et al., 2009; Martinez-Huitle
and Brillas, 2009; Nidheesh and Gandhimathi, 2012; Sires and Brillas, 2012).
Among these indirect electrooxidation methods, the most popular technique
is the EF process in which •OH radicals are produced in the electrochemi-
cally assisted Fenton reaction (reaction 1) involving in situ electrogenerated
H2 O2 and electroregenerated Fe2+ (Fenton’s reagent) (Oturan, 2000; Brillas
et al., 2009).
In addition, •OH radicals can be anodically electrogenerated by water
oxidation in variable amounts, according to the nature of the anode material.
The EF process can be conducted either in divided or in undivided electro-
chemical cells. In the latter case, it can take advantage of the oxidation reac-
tions arising from the simultaneous production of both anode and cathode,
which is more efficient than the classical, above-described AO process for de-
struction of organic pollutants. In order to enhance the EF process efficiency,
its coupling has been recently proposed with other AOPs, such as PEF, solar
photo-electro-Fenton, sonoelectro-Fenton, and peroxi-coagulation.
5.2.1 PRINCIPLE AND FUNDAMENTALS OF EF PROCESS
The EF process is an indirect EAOPs since hydroxyl radicals are generated
via the Fenton’s reagent (mixture of H2 O2 and ferrous iron ions) and through
the Fenton reaction (reaction 1) in homogeneous medium, including the in
situ electrogeneration of H2 O2 and electroregeneration of Fe2+ ions that con-
stitute Fenton’s reagent. H2 O2 is in situ electrogenerated by a two-electron
reduction of dissolved O2 in acidic medium (reaction (31)) in presence of a
catalytic amount of ferrous ions (Oturan et al., 2000; Brillas et al., 2009).

O2 +2H+ +2e− → H2 O2 (31)

 TABLE 2. Selected examples of oxidative degradation/mineralization of organic pollutants in water by anodic oxidation process
Pollutant Expérimental conditions Matrix Results obtained References
Pesticides
Chlorophenoxy Degradation of chlorophenoxyacid herbicides 4-CPA, Millipore Milli-Q system Total mineralization of solutions of pH 3.0 containing Brillas et al. (2004)
herbicides MCPA, 2,4-D, and 2,4,5-T in aqueous medium of pH pure water 100 mg L−1 of TOC of chlorophenoxy herbicides
3.0 using BDD anode (3 cm2) and 3.1 cm2 4-CPA, MCPA, 2,4-D, and 2,4,5-T was reached even
O2 -diffusion carbon-PTFE cathode; I = 100–450 mA at low current values when using BDD anode while
mineralization efficiency remain low with Pt anode.
Amitrole Comparative oxidation of amitrol by AO on BDD anode Deionized water Faster degradation with BDD anode than with a Da Pozzo et al.
and EF process at pH 3. platinum anode. Amitrole decay follows pseudo (2005)
first-order reaction. NO3 − and NH4 + are released to
the medium during mineralization.
Parathion Use of BDD anode both as sensor (analytical Natural water Detection and quantification limits were determined as Pedrosa et al.
application) and anode in its oxidation 2.4 and 7.9 μg L−1, respectively. When BDD is used (2006)
(environmental application. In the last case a high as anode for depollution, progressive diminution of
voltage of 3 V was applied. solution TOC was observed indication conversion of
parathion to CO2 and H2 O.
Mecoprop 50 cm2 Ti/PbO2 or BDD electrodes were used as anode, Milli-Q pure water Oxidation power of PbO2 and BDD anodes was Sirés et al. (2008)
the cathode being a 50 cm2 stainless steel. Currents compared in mineralization of mecoprop. BDD
of 1–3 A were applied to treat the solutions anode exhibits faster oxidation rate and greater
containing Initial mecoprop concentration from current efficiencies compared to PbO2 . GC–MS
178 mg L−1 to 700 mg L−1. analyses revealed that the electrochemical
degradation pathway involves the same aromatic and
carboxylic acid intermediates when using PbO2 or
BDD anodes and that oxalic acid is the ultimate
by-product prior to conversion to CO2 .
Propham BDD thin film anode (3 × 4 cm), Pt gauze cathode, pH: Deionized water Propham decay follows pseudo-first-order kinetics with Özcan, et al.
3–11, applied current: 30–500 mA, supporting kapp = 4.8 × 10−4 s−1 at 100 mA and 35◦ C in the (2008b)
electrolyte: Na2 SO4 50 mM presence of 50 mM Na2 SO4 in acidic medium of pH:
3. Temperature of 35 ◦ C was found as optimal
temperature. Propham degradation is fast in NaCl
medium due to the mediated oxidation by formed
HOCl. Hydroquinone, cathecol, benzoquinone,
p-hydroxyphenyl carbamic acid isopropyl ester, and
o-hydroxyphenyl carbamic acid isopropyl ester were
found as oxidation reaction intermediates.
(Continued on next page)

2601
 TABLE 2. Selected examples of oxidative degradation/mineralization of organic pollutants in water by anodic oxidation process (Continued)
Pollutant Expérimental conditions Matrix Results obtained References

2602
Methamidophos Experiments were performed using a divided cell Deionized water The oxidation power of Pb/PbO2 , Ti/SnO2 , and Martı́nez-Huitle
with compartment of 100 mL capacity, at pH 2, Si/BDD anodes was compared. Results showed et al. (2008)
with Na2 SO4 as supporting electrolyte that the electrode efficiency is dependent on the
applied current density. Si/BDD electrode
showed better efficiency, reaching almost total
mineralization with the application of current
density of 50 mA cm−2.
Chlorpyrifos Electrolyses were conducted in a two compartments Double distilled water The kinetic mineralization was evaluated by means Samet et al. (2010)
and thermostated cell under galvanostatic of the COD measurement which is enhanced by
conditions using Nb/PbO2 anode and graphite increasing applied current density and
carbon bar as cathode and applying temperature. The best COD removal of 76% was
10–50 mA cm−2 in 0.1 M HClO4 as supporting obtained when using an current density of
electrolyte. 50 mA cm−2, initial COD = 450 mg O2 L−1 and
at 70 ◦ C in 10 hr electrolysis time
Atrazine 30 mg L−I atrazine solutions were treated by AO Ultra-pure water with High mineralization rates were obtained by different Borras et al.
using BDD anode and stainless steel or O2 resistivity > 18 M cm electrochemical methods using BDD anode. It (2010)
2
diffusion cathode (3 cm ) able to generate H2 O2 . was observed that mineralization rate is limited
Electrolyses were conducted in open and by the oxidation of persistent by-products formed
undivided cylindrical cell containing 100 mL during oxidation of atrazine. Optimum conditions
solution at pH 3. of 300 mA and pH 3.0 have been determined for
the treatment. PEF process using BDD anode
exhibited higher oxidation power than AO with
H2 O2 generation, reaching finally 95%
mineralization.
Pharmaceuticals
Paracetamol Graphite bar as cathode and BDD/Pt as anode, Millipore Milli-Q water Mineralization process accompanied with release of Brillas et al. (2005)
0.05 M Na2 SO4 as SEC, at pH = 2.0–12.0 and NH4 + and NO3 −; the current efficiency increased
25–45◦ C, with paracetamol < 1 g L−1 with raising drug concentration and temperature;
oxalic and oxamic acids were detected as ultimate
products, completely removed with Pt and its
kinetics followed a pseudo-first-order reaction
with a constant rate independent of pH.
Salicylic acid Solutions containing 164 mg L−1 salicylic acid of pH Millipore Almost total mineralization was reached with BDD Guinea et al.
3.0 have been degraded by AO with Milli-Q water anode. Salicylic acid decay followed a (2008)
electrogenerated H2 O2 at the gas diffusion pseudo-first-order kinetics. 2,3-Dihydroxybenzoic,
cathode. 2,5-dihydroxybenzoic, 2,6-dihydroxybenzoic,
α-ketoglutaric, glycolic, glyoxylic, maleic,
fumaric, malic, tartronic, and oxalic acids are
detected as oxidation products.
 Enrofloxacine Solutions of veterinary antibiotic enrofloxacin in Millipore AO-H2 O2 with BDD yielded the poorest Guinea et al.
(antibiotic) 0.05 M Na2 SO4 of pH 3.0 were comparatively Milli-Q water mineralization because the limitation of the (2010)
degraded by different electrochemical AOPs, oxidation process to the anode surface (mass
including AO with electrogenerated H2 O2 on transport limitation) in contrast of PEF or solar
BDD anode. PEF in which oxidation supplementary oxidation
occurs in the solution bulk. Enrofloxacin decay
always followed pseudo-first-order reaction.
Primary aromatic by-products and short
intermediates including polyols, ketones,
carboxylic acids, and N -derivatives were detected
by GC–MS analysis.
Beta-blockers Oxidative treatments of 10 L solutions with Top water Mineralization rate of 88–93% were obtained. Decay Isarain-Chavez
100 mg L−1 of TOC of beta-blockers atenolol, kinetics of beta-blockers followed pseudo et al. (2011)
metoprolol tartrate (2:1) and propranolol first-order reaction kinetics. Decay kinetics was
hydrochloride was conducted in a recirculation found accelerated by additional production of
•
flow plant equipped with Pt and/or BDD anode OH from the action of UV light (PEF) or solar
and gas diffusion cathode, at pH 3 and 35 ◦ C. irradiation (solar PEF). Bet-blockers and all
aromatic intermediates were destroyed by
hydroxyl radicals at the end of treatment.
Ultimate carboxylic acids like oxalic and oxamic
remained in the treated solutions in AO, but they
are mineralized when using solar PEF.
Sulfanilic acid AO of sulfanilic acid solutions of pH interval 2.0–6.0 Millipore Overall mineralization was achieved under all El-Ghenymy et al.
(antibiotic) was studied in divided and undivided cells with a Milli-Q water experimental conditions tested due to the (2012)
BDD anode and a stainless steel cathode. efficient destruction of sulfanilic acid and all its
by-products with hydroxyl radicals generated at
the BDD anode from water oxidation. Decay of
sulfanilic acid followed pseudo-first-order
kinetics. Hydroquinone and p-benzoquinone
were identified as aromatic intermediates by
GC-MS analysis. Maleic, acetic, formic, oxalic, and
oxamic acids were detected as generated
carboxylic acids.
4-Chlorophenol Oxidation of 4-chlorophenol was investigated on The experimental results have been compared with Rodrigo, et al.
synthetic diamond film electrodes in sulfuric a theoretical model involving a fast oxidation of (2001)
acidic medium. Experiments were carried out in 4-chlorophenol to p-benzoquinone through the
three-electrode cell equipped with BDD anode, formation of phenoxy hexadienyl radical. Good
Pt counter electrode, and Hg/Hg2 SO4 ·K2 SO4 agreement between experimental data and
reference electrode. theoretical model. The p-benzoquinone was
detected as main aromatic intermediate.
Carboxylic acids such as maleic, formic, and
oxalic acids were found as main by-products.
(Continued on next page)

2603
 TABLE 2. Selected examples of oxidative degradation/mineralization of organic pollutants in water by anodic oxidation process (Continued)
Pollutant Expérimental conditions Matrix Results obtained References

2604
Synthetic dyes
Acid Blue 12 Electrochemical oxidation of synthetic wastewater Synthetic wastewater It was found that AO with BDD anode is suitable Panizza and
containing acid blue 22 on a BDD electrode was containing Acid Blue for completely removing COD and effectively Cerisola (2008)
studied, using cyclic voltammetry and bulk 22 decolorizing synthetic wastewaters under optimal
electrolysis. experimental conditions of flow rates (i.e.
300 dm3 h−1) and current density (i.e.
20 mA cm−2). 97% of COD was removed in 12 hr
electrolysis involving energy consumption of
70 kWh m−3.
Acid Orange 7 Degradation of dye Acid Orange 7 was conducted Deionized water The absolute rate constant of the AO 7 Hammami et al.
comparatively in acidic medium of pH 3.0 using hydroxylation reaction was determined as (2008)
Pt and BDD anodes and carbon-felt cathode. (1.10 ± 0.04) × 1010 M−1 s−1 by using
competition kinetic method. High mineralization
ratio of 98% in terms of TOC removal was
obtained after 9 hr of electrolysis at 250 mA. The
follow-up of the solution toxicity showed the
formation of intermediates more toxic than AO 7
and the connection between toxicity and
aromaticity. A mineralization reaction pathway of
AO 7 by hydroxyl radicals was proposed.
Anionic Synthetic solution of sodium dodecyl benzene Synthetic SDBS solution In the case Ti–Ru–Sn ternary oxide anode, the Panizza et al.
surfactant sulfonate (SDBS) and a real car wash wastewater and real car wash complete removal of COD and SDBS was (2005)
were treated by AO at 75 mA cm−2; using a wastewater obtained only in the presence of chloride ions.
Ti–Ru–Sn ternary oxide and BDD anode. The Chlorine-mediated oxidation at the Ti–Ru–Sn
BDD or TiRuSnO2 anodes and the stainless-steel ternary oxide anode allowed a faster COD
cathode were square with a geometric area of removal of both the synthetic solution and real
25 cm2. car wash wastewater. In the case of BDD anode,
the mineralization of the sodium dodecyl
benzene sulfonate was achieved in all
experimental conditions
Synthetic Electrolyses were carried out in galvanostatic Synthetic vegetable Experimental results showed that both the Panizza and
tannery effluent conditions using Ti/PbO2 (25 cm2) and Ti/TiRuO2 tannery wastewater electrodes performed complete mineralization of Cerisola (2004)
2
(25 cm ) anodes under different experimental the wastewater; the oxidation took place on the
conditions. PbO2 anode by direct electron transfer and
indirect oxidation mediated by active chlorine,
while it occurred on the Ti/TiRuO2 anode only by
indirect oxidation. Therefore, the Ti/TiRuO2
required almost same energy consumption for
complete COD removal; it was more stable and
did not release toxic ions, so it seems to be the
best candidate for industrial applications.
 Advanced Oxidation Processes in Water/Wastewater Treatment 2605

The first cathode material used in H2 O2 generation was mercury pool

(Hg) (Tomat and Vecchi, 1971; Tzedakis et al., 1989; Oturan et al., 1992;
Oturan and Pinson, 1995), but it has been discarded due to Hg potential
toxicity, when the method was applied to wastewater treatment. Therefore,
Hg has been replaced by various carbonaceous materials, including graphite
(Sudoh et al., 1986), carbon-PTFE O2 diffusion (Brillas et al., 1995; Harring-
ton and Pletcher, 1999), and three-dimensional electrodes such as carbon felt
(Oturan et al., 2000, 2009a), activated carbon fiber (Wang et al., 2005), retic-
ulated vitreous carbon (Alvarez-Gallegos and Pletcher, 1998), carbon sponge
(Özcan et al., 2008a), and carbon nanotubes (Fu et al., 2007). The use of
carbon cathodes is of interest because, unlike mercury, carbon is nontoxic
and exhibits a high overpotential for H2 evolution (good current efficiency)
and low catalytic activity for H2 O2 decomposition. Carbon material has also
good stability, conductivity, and chemical resistance.
Although H2 O2 is a weak oxidant with E ◦ (O2 /H2 O2 ) = 1.78 V/SHE)
compared to other chemical oxidants, such as O3 (E ◦ (O3 /O2 ) = 207 V/SHE),
its oxidation power is significantly enhanced in presence of Fe2+ ions Fen-
ton’s reaction, which can be again attributed to the formation of strongly
oxidant hydroxyl radical via the reaction (1). Fe3+ formed by Fenton reac-
tion is then reduced electrochemically at the cathode to regenerate ferrous
iron (reaction (34)) in order to catalyze the •OH Fenton’s reaction.

Fe3+ +e− → Fe2+ (32)

Simultaneous generation of H2 O2 and regeneration of ferrous iron at the

cathode allow one to form •OH radicals in a continuous and catalytic man-
ner. Then, formed •OH radicals rapidly react in the bulk with organic pollu-
tants either by hydrogen atom abstraction (reaction (33)) or by OH addition
(reaction (34)):

Organic pollutants +• OH → oxidation intermediates (33)

Intermediates + → → → CO2 + H2 O + inorganic ions (34)

Three-dimensional electrodes, particularly carbon-felt cathodes, provide high

efficiency in water/wastewater treatment by EF process, since these elec-
trodes present a high surface-to-volume ratio, combined with a low cost
and easy handling. Thanks to particular hydrodynamic conditions, these
three-dimensional cathodes have large specific area and large mass transfer
coefficients of dissolved O2 , which increases the rates of generation of the
Fenton’s reagent, of formation of •OH radicals via the Fenton’s reaction,
and of degradation of organic pollutants (Kesraoui-Abdessalem et al., 2008;
Oturan et al., 2010a; Panizza and Oturan, 2011).
The EF process presents several major advantages relative to the classi-
cal chemical Fenton reaction (Harrington and Pletcher, 1999; Oturan, 2000;
2606 M. A. Oturan and J.-J. Aaron

FIGURE 5. (a) Sketch of an open and stirred two-electrode undivided, bench-scale tank
reactor with a 60 cm2 carbon-felt cathode fed with compressed air for treatment of organic
containing solutions by EF process and, (b) Schematic representation of the main reactions
involved in the EF process. RH denotes an unsaturated compound that undergoes dehy-
drogenation, while Ar denotes an aromatic compound that is hydroxylated. Adapted with
permission from Brillas et al. (2009) and Oturan et al. (2008a).

Oturan and Brillas, 2007; Brillas et al., 2009): (i) in-site production of H2 O2
avoiding the risks related to its transport, storage, and handling, (ii) possibil-
ity of controlling degradation kinetics and performing mechanistic studies,
(iii) higher removal rate of organic pollutants due to Fe2+ continuous regen-
eration at the cathode, (iv) no need of chemical reagents and no formation
of sludge, and (iv) feasibility of overall mineralization at a relatively low cost
by optimizing the operation parameters.
Figure 5 shows schemes of an undivided tank reactor for production
•
of OH radicals (a) and of the main reactions involved in the EF process
(b) (Oturan et al., 2008; Brillas et al., 2009):

5.2.2 IMPLEMENTATION AND DEVELOPMENT

The early works dealing with in situ formation of Fenton’s reagent by simul-
taneous cathodic reduction of dioxygen and ferric iron ions were carried out
by Tomat et al. (Tomat and Vecchi, 1971; Tomat and Rigo, 1976, 1980). These
authors reported the one-step oxidation of organic compounds, including
benzene, toluene, and cyclohexane, into the corresponding hydroxylated
derivatives in H2 SO4 medium and on Hg pool as cathode. However, these
reactions were performed with difficulty, and weak yields were obtained,
due to the nonselectivity of •OH radicals. Then, Tzedakis et al. (1989) ob-
tained a better conversion yield of 70% for the EF reaction of benzene into
phenol by using a batch reactor in continuous mode with an Hg cathode.
 Advanced Oxidation Processes in Water/Wastewater Treatment 2607

More recently, Oturan et al. have carried out the polyhydroxylation of sali-
cylic acid (Oturan et al., 1992) and benzoic acid (Oturan and Pinson, 1995),
as well as that of some chlorophenoxy herbicides and other aromatic pesti-
cides (Oturan et al., 1999a) at pH 3.0, by means of •OH radicals generated
on a Hg cathode (applied potential E cat = –0.5 V/SCE), and with addition
of Fe3+ as catalyst. Also, efforts have been made for electro-generating H2 O2
with environmental-friendly electrodes built in carbonaceous materials, in
order to avoid the Hg cathode toxicity. For instance, the Oturan’s group
has reported the production of monohydroxylated metabolites of the drug
riluzole (Oturan et al., 1999b), and of mono-, di-, and trihydroxylated deriva-
tives of chlorophenoxy acid pesticides (Aaron and Oturan, 2001) by EF with
a carbon-felt cathode at pH 2.0.
Sudoh et al. (1986) performed the first application of electrogener-
ated Fenton’s reagent to wastewater treatment. These authors realized the
degradation of O2 -saturated phenol solutions with Fe2+ as catalyst in a two-
compartment cell (E cat = −0.6 V vs. Ag/AgCl/KCl (satd) cathode). Following
this study, it appeared an increasing number of papers on the destruction
of various toxic and refractory organic pollutants in water by means of EF
and related processes, particularly the extensive, remarkable work of the
Brillas and Oturan research groups, who used, respectively, carbon-PTFE
O2 -diffusion and carbon-felt cathodes (Oturan, 2000; Oturan et al., 2000,
2009a; Brillas et al., 2009).
When a low O2 -overvoltage anode like Pt was utilized in a one-
compartment cell, the oxidation of water into O2 (reaction (35)) simply
occurred. In this case, O2 needed for production of H2 O2 (reaction (31)) was
generated in an anodic reaction (Oturan et al., 2001), and, consequently, the
EF process constituted an overall catalytic system. As can be seen in Figure 6,
two catalytical cycles took place in this system, continuously regenerating
Fe2+ and H2 O2 and producing •OH radicals via the Fenton reaction.

2H2 O → O2 + 4H+ + 4e− (35)

A significant enhancement of EF process has been achieved by replac-

ing the classical Pt anode with the emerging BDD anode. In this case, the
EF process became more potent because of the simultaneous generation of
supplementary heterogeneous •OH radicals, BDD(•OH), at the anode sur-
face (reaction (25)) in addition to those produced in bulk solution (reaction
(1)). Moreover, the use of BDD anode in the treatment provided two other
advantages: (i) oxidizing power of BDD(•OH) higher than other anodes due
to a larger O2 overvoltage and (ii) high oxidation window of BDD anode
(about 2.5 V), which allowed to directly oxidize organic pollutants (Brillas
and Martinez-Huitle, 2011).
An outstanding, very recent example of the EF process improvement
with BDD anode concerns the mineralization of the very refractory and
2608 M. A. Oturan and J.-J. Aaron

FIGURE 6. Schematic representation of the electrocatalytic production of hydroxyl radicals

by the EF process with a low O2 -overvoltage anode. (Source: Oturan et al., 2001)
C Elsevier. Reproduced by permission of Elsevier. Permission to reuse must be obtained from
the rightsholder.

toxic herbicide atrazine in aqueous medium (Oturan et al., 2012). Indeed,

a large variety of AOPs, including chemical, photochemical, and photocat-
alytic processes, have been already applied to remove atrazine from aqueous
medium, but these AOPs produced only the persistent end product cya-
nuric acid (2,4,6-trihydroxy-1,3,5-triazine) as predominant by-product (De
Laat et al., 1999), with 40–60% mineralization yields, corresponding to the
atrazine side-chain mineralization. In the case of other, previously applied
EAOPs, the atrazine mineralization attempts led to the same end product
(Malpass et al., 2006; Mamián et al., 2009). For example, the atrazine AO
with a DSA anode removed the solution TOC by only 46% (Sun and Pig-
natello, 1993). In contrast, the use of a BDD anode in the EF process allowed
one to obtain an almost total mineralization (97% TOC removal) of atrazine
aqueous solutions (Figure 7) (Oturan et al., 2012). Comparative TOC removal
curves clearly showed the greater mineralization power of EF-BDD anode
relative to the classical EF-Pt process in the treatment of a very recalcitrant
pollutant like atrazine (Figure 7). As already mentioned, this great miner-
alization power can be ascribed to the simultaneous generation of large
amounts of heterogeneous BDD (•OH) radicals at the anode surface and of
homogeneous •OH radicals in the solution bulk.
Furthermore, the same authors demonstrated that cyanuric acid, which
was a long time considered as the stable end product of atrazine mineral-
ization, and considered as refractory to the action of •OH radicals, can be
attacked and even mineralized by •OH generated in the EF-BDD process
(Figure 8) (Oturan et al., 2012). Indeed, as can be seen, a very high min-
eralization degree as 90% TOC removal was reached during the EF-BDD
treatment of cyanuric acid aqueous solutions. This result again highlighted
 Advanced Oxidation Processes in Water/Wastewater Treatment 2609

10
EF-Pt
8 AO-BDD

-1
EF-BDD

TOC / mg L
6

0
0 2 4 6 8
Electrolysis time / h

FIGURE 7. TOC removal kinetics during the mineralization experiment of 0.1 mM atrazine
solution by EF process with Pt and BDD and carbon-felt cathode. EF-Pt: EF process with
Pt anode, AO-BDD: anodic oxidation with BDD anode and EF-BDD: EF process with BDD
anode. (Source: Oturan et al., 2012)
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from the rightsholder.

the strong mineralization power of the EF-BDD process, since the EF pro-
cess with Pt anode led to only 4% of TOC removal, which confirms the great
interest of using a BDD anode in EF process for the treatment of POPs.

5.2.3 EFFECT OF OPERATING PARAMETERS ON THE EF PROCESS EFFICIENCY

A number of operating parameters, such as solution pH, applied current,
catalyst concentration, medium (supporting electrolyte), and initial organic
content, were found to affect the EF process effectiveness.

FIGURE 8. Mineralization of cyanuric acid (C 0 = 25.8 mg L−1) aqueous solution by anodic

oxidation with BDD anode (AO-BDD), and EF with Pt (EF-Pt) and BDD (EF-BDD) anodes at
1 A constant current. (Source: Oturan et al., 2012)
C Springer. Reproduced by permission of Springer. Permission to reuse must be obtained
from the rightsholder.
2610 M. A. Oturan and J.-J. Aaron

The solution pH constitutes one of the key parameters of the EF pro-

cesses. First, an optimal value of pH 2.8 was reported by Sun and Pignatello
(1993) for the Fenton’s and Fenton-related processes, and then was con-
firmed by other authors (Boye et al., 2002a; Diagne et al., 2007). Neverthe-
less, several studies have indicated that the EF process might be carried out
more or less effectively in the pH range of 2.5–3.5. For pH values >4, the iron
ions (catalyst) are precipitated under the Fe(OH)3 form, whereas in the case
of pH values <2.5, the Fenton process efficiency decreases because of the
formation of peroxonium ions (H3 O2 +), which makes the H2 O2 electrophilic
and reduces its reactivity towards Fe2+ (Feng et al., 2003).
It has been also shown that lower pH values transformed •OH radicals
into HO2 • radicals according to reaction (36), which decreased the Fenton
process efficiency (Tang and Huang, 1996).

H2 O2 +• OH → HO•2 + H2 O (36)

Applied current (or current density) is another crucial parameter, since it

governs the formation rate of H2 O2 (reactions (25) and (31)) as well as the
regeneration rate of ferrous iron (reaction (32)), and consequently the gen-
eration rate of OH via Fenton reaction. In general, the EF process efficiency
was found to increase with the applied current until a definite, limit value for
which parasitic reactions could not be considered as negligible. Therefore,
in the case of applied current values larger than this limit value, the current
efficiency (or mineralization current efficiency) decreased because of the en-
hancement of wasting reactions, such as H2 evolution at the cathode or •OH
coupling at the anode (case of the BDD anode). The optimal current value
could be experimentally determined before each test for degradation kinetics
and/or mineralization efficiency, as shown in the case of the herbicide piclo-
ram degradation (Figure 9) (Özcan et al., 2008c). The total disappearance of
picloram (0.125 mM) was reached within less than 5 min for applied current
values of 200, 300, and 500 mA. Therefore, it seemed convenient to choose
an applied current of 200 mA as an optimal value, which minimized the
energy consumption for practically the same efficiency and reaction time.
The third important operating parameter is the catalyst concentration,
particularly when ferrous (Fe2+) or ferric (Fe3+) iron is used as catalyst. As
an example, the effect of Fe3+ concentration (catalyst) on the antimicrobial
chlorophene decay during the EF treatment is displayed in Figure 10 (Sirés
et al., 2007b). It can be seen that the chlorophene degradation rate increased
when the catalyst (Fe3+) concentration decreased from 2.0 to 0.1 mM. Indeed,
the time needed for complete removal of chlorophene (50 mg L−1) was about
four fold higher for the 2 mM Fe3+ solution than for the 0.1 mM Fe3+ one.
This behavior can be explained by the enhancement of the parasitic reaction
occurring between Fe2+ and •OH when increasing the catalyst concentration
 Advanced Oxidation Processes in Water/Wastewater Treatment 2611

FIGURE 9. Effect of applied current on degradation kinetics of 150 mL picloram (C 0 :

0.125 mM) solution at pH 3 by EF process: (x): 30 mA, (): 60 mA, (ж): 100 mA, (♦):
200 mA, (): 300 mA and (+) 500 mA. [Fe3+]: 0.1 mM, [Na2 SO4 ]: 50 mM. (Source: Özcan
et al., 2008)
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the rightsholder.

(reaction (37)).

Fe2+ +• OH → Fe3+ +OH− (37)

When Fe2+ was used as catalyst, reaction (37) became competitive with the
degradation reaction of organic pollutant with •OH for high concentrations.

60

50 (b)
[Chlorophene] / mg L-1

40

30

20

10

0
0 5 10 / min 15
Time 20 25

FIGURE 10. Influence of Fe3+ concentration (as catalyst) on chlorophene decay during the
EF treatment of 200 mL of 50 mg L−1 solutions at pH 3.0 and 300 mA using a Pt/carbon-felt
cell. [Fe3+]0 : () 0.1 mM, () 0.2 mM, () 0.5 mM, () 1.0 mM, () 2.0 mM. (Source: Sirés
et al., 2007)
C Elsevier. Reproduced by permission of Elsevier. Permission to reuse must be obtained from
the rightsholder.
2612 M. A. Oturan and J.-J. Aaron

When the catalyst was Fe3+, it was rapidly reduced into Fe2+ at the cathode,
and, in these conditions, a higher Fe3+ initial concentration resulted in a
higher Fe2+ concentration in the solution, leading to a decrease of the EF
process efficiency.

5.2.4 APPLICATION OF THE EF PROCESS TO THE TREATMENT OF ORGANIC

POLLUTANTS
Since its first applications to the treatment of wastewaters which occurred
more than a decade ago (Oturan et al., 1999a, 2009b; Oturan, 2000), the
EF process has considerably developed for the decontamination of waters
containing toxic and POPs, such as pesticides, industrial pollutants, synthetic
dyes, pharmaceuticals, personal care products, etc.
Regarding pesticides, a pioneering work (Oturan, 2000; Oturan et al.,
2009b) has demonstrated that the EF process allowed one to rapidly and al-
most totally mineralize (>95% TOC decay) a 1 mM 2,4-dichlorophenoxyacetic
acid (2,4-D) aqueous solution in the presence of 1 mM Fe3+, using an un-
divided cell equipped with carbon-felt cathode and Pt anode (Figure 11a).
The decay kinetics of the initial chlorophenoxy herbicide was fast, attaining
total removal after the passage of only a 400 C charge. Oxidation of 2,4-D by
•
OH generated in the process led to the formation of hydroxyl derivatives,
such as 2,4-dichlorophenol, 2,4-dichlororesorcinol, 4,6-dichlororesorcinol, 2-
chlorohydroquinone, and 1,2,4-trihydroxybenzene. As shown in Figure 11b,
the first, three dichloro-phenol derivatives disappeared after consuming 700
C, while 2,4-D was destroyed with a charge of 400 C. These results con-
firmed that the attack of •OH on the 2,4-D aromatic ring led to the formation
of its hydroxyl derivatives, along with other processes involving dechlori-
nation and dehydrogenation. The TOC progressive decrease indicated that
the formed intermediates reacted with •OH similarly to 2,4-D, and yielded
poly-hydroxylated products and quinones, before undergoing an oxidative
ring opening reaction to generate short-chain carboxylic acids. Another work
(Oturan et al., 1999a) confirmed the fast oxidation kinetics of chlorophenoxy
acid herbicides, namely mecoprop (2-(4-chloro-2-methylphenoxy)propionic
acid), CPMP (2-(4-chlorophenoxy)-2-methylpropionic acid), 2,4-DP (2-(2,4-
dichlorophenoxy)propionic acid), and 2,4,5-T (2,4,5-trichlorophenoxyacetic
acid). In this study, Oturan et al. (1999a) reported the formation of polyhy-
droxyphenols and quinones in a first step, and the complete destruction of
the aromatic and/or cyclic derivatives upon exhaustive electrolysis.
This behavior has been also observed for a polychlorinated pesticide,
such as PCP (pentachlorophenol) (Oturan et al., 2001) in which •OH rad-
icals were able to perform an ipso-attack on chlorine positions to form
hydroxyl intermediates. Then, these intermediates were quickly transformed
into dechlorinated by-products, including tetrachloro-o-benzoquinone and
tetrachloro-p-benzoquinone. A TOC abatement (mineralization) of 82% was
 Advanced Oxidation Processes in Water/Wastewater Treatment 2613

100
(a)
80

-1
TOC / mg L
60

40

20

0
0 500 1000 1500 2000
Charge / C

0.20
2.0
(b)
Concentration / mM

1.5
0.15
Concentration / mM

1.0

0.5

0.10 0.0
0 200 400 600 800
Charge / C

0.05

0.00
0 200 400 600 800
Charge / C

FIGURE 11. 1. Electro-Fenton treatment of 2,4-D (2,4-dichlorophenoxy)acetic acid herbicide.

(a) Time-course of mineralization of solution of 1 mM 2,4-D (in terms of TOC removal), and
(b) Evolution of the concentration of 2,4-D (inset) and three of hydroxylated derivatives:
2,4-dichlorophenol (), 2,4-dichlororesorcinol (), and 4,6-dichlororesorcinol () as function
of electrolysis time. (Source: Oturan, 2000).
C Springer. Reproduced by permission of Springer. Permission to reuse must be obtained
from the rightsholder.

found after consumption of 1500 C, while the total release of all chlorine as
chloride ions was achieved for a charge of 600 C. It was demonstrated that
the affinity of chlorophenols for hydroxyl radicals depended on the num-
ber of aromatic chlorine substituents, the rate constant values decreasing
when the number of chlorine substituents increased (Oturan et al., 2009a).
Thereafter, the good efficiency of the EF process has been demonstrated for
various families of herbicides, namely organophosphorus (Guivarche et al.,
2003a; Diagne et al., 2007), imidazolines (Kaichouh et al., 2004), phenylurea
(Edelahi et al., 2003; Kesraoui-Abdessalem et al., 2008; Oturan et al., 2010a,
2011), triazines (Oturan et al., 2009b, 2012; Balci et al., 2009b), phosphonates
2614 M. A. Oturan and J.-J. Aaron

(Balci et al., 2009a), triketones (Murati et al., 2012), as well as insecticides

(Oturan et al., 2010b), fungicides (Özcan et al., 2008b), and mixture of pesti-
cides (Oturan and Oturan, 2005; Kesraoui-Abdessalem et al., 2010a, 2010b).
Moreover, the Brillas’ team has developed an EF treatment, using a
carbon-PTFE gas diffusion cathode, to eliminate pesticides from aqueous so-
lutions (Boye et al., 2002a, 2003b; Brillas et al., 2003b; Borras et al., 2011), and
to almost totally mineralize the organic content of these solutions. Also, the
EF process has been applied by several authors to efficiently remove organic
pollutants from several kinds of wastewaters and/or effluents from different
origins, including industrial pollutants (Brillas et al., 1998; Oturan et al., 2000;
Gözmen et al., 2003; Pimentel et al., 2008), industrial wastewaters (Panizza
and Cerisola, 2001), olive oil mill wastewater (Bellakhal et al., 2006; Zhang
et al., 2006; Wang et al., 2012), and reverse osmosis concentrates (Trabelsi
et al., 2012; Zhou et al., 2012). In addition, a large number of studies has
been focused on the oxidation/mineralization of synthetic dyes (Guivarch
and Oturan, 2004; Wang et al., 2005, 2008; Hammami et al., 2007; Guivarch
et al., 2003b; Oturan et al., 2008; Khataee et al., 2009; Özcan et al., 2009;
Panizza and Cerisola, 2009; Garcia-Segura et al., 2011b; Panizza and Oturan,
2011; Bouafia-Chergui et al., 2012; Hammami et al., 2012; Khalfaoui et al.,
2012). These works concerned the decolorization of dyes, determination of
decay kinetics and reaction rate constants with hydroxyl radicals, mineral-
ization yields, and elucidation of plausible mineralization pathways, based
on the identification of aromatic/cyclic reaction intermediates, short-chain
carboxylic acids and inorganic end products.
Recently, important efforts have been devoted to studies on the re-
moval of a new class of emerging pollutants, including pharmaceuticals and
personal care products, because of their occurrence in natural waters and
their potentially toxic effects on aquatic species (Sires and Brillas, 2012).
The group of Brillas has realized a pioneering work on the paracetamol
mineralization in aqueous medium (Sires et al., 2004), followed by several
works on antimicrobials (Sirés et al., 2007c; Skoumal et al., 2008; Oturan
et al., 2009b), nonsteroidal antiinflammatory drugs (Guinea et al., 2008; Sk-
oumal et al., 2009; Brillas et al., 2010; Loaiza-Ambuludi et al., 2013), and
beta-blockers (Isarain-Chavez et al., 2010a, 2010b; Sirés et al., 2010), with
high treatment efficiency, including the determination of kinetics, mechanis-
tic assessments, and degradation pathways. Also, Dirany et al. (2010, 2011,
2012) and El-Ghenymy et al. (2012) have reported detailed investigations
on water contaminated by sulfonamide antibiotics, in which the evolution of
toxicity of treated solutions was monitored, using the MicrotoxR biolumines-
cence method. The toxicity changes observed during the oxidative treatment
of a sulfamethoxazole (SMX) synthetic aqueous solution by the EF process
revealed an interesting behavior (Figure 12) (Dirany et al., 2011).
Indeed, the different Vibrio fischeri bacteria luminescence inhibition
peaks appearing as a function of EF treatment time can be related to the
 Advanced Oxidation Processes in Water/Wastewater Treatment 2615

120,00
(a) 300 mA
120 mA
90,00
60 mA

% Inihb 15 min
30 mA
60,00

30,00

0,00
0 30 60 90 120 150 180 210 240

-30,00

Time / min

120,00
AMI
(b) BZQ
90,00
% Inh 15 min

60,00

30,00

0,00
0 30 60 90 120 150 180 210 240 270
-30,00

Time / min

FIGURE 12. Curves of evolution of the Vibrio fischeri bacteria luminescence inhibition with
time during the EF treatment of (a) SMX aqueous solutions, and (b) its cyclic derivatives AMI
and BZQ diluted aqueous solutions, after an exposure time of 15 min. V = 250 mL. [SMX]0 =
0,208 mM, [AMI]0 = 0.016 mM; [BZQ]0 = 0.018 mM. [Fe2+] = 0.2 mM. [Na2 SO4 ] = 50 mM.
pH = 3. I = 30, 60, 120, 300 mA for (a), and I = 60 mA for (b) Pt anode and carbon-felt
cathode. (Source: Dirany et al., 2011)
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from the rightsholder.

formation of primary, then secondary or tertiary aromatic (and/or cyclic)

products which can be more toxic than the parent compound. The high-
performance liquid chromatography (HPLC) results of this study showed
that 3-amino-5-methylisoxazole (AMI) and p-benzoquinone (BZQ) were the
primary oxidation products of SMX. Figure 12b was obtained by using di-
luted solutions of AMI and BZQ, effectively attained during the EF oxidation
of SMX (Dirany et al., 2011). The MicrotoxR bioluminescence method al-
lowed us to demonstrate that both aromatic (and/or cyclic) intermediates
were at least partly responsible of the evolution of toxicity of SMX solutions
during the EF treatment. Moreover, these results indicated that at least one
2616 M. A. Oturan and J.-J. Aaron

of the formed degradation aromatic (and/or cyclic) intermediates, namely

BZQ, was clearly more toxic than SMX itself towards the luminescent bacte-
ria Vibrio fischeri. As can be seen in Figure 12a, it occurred a strong initial
increase of the luminescence inhibition, with a maximum lasting between
about 10 min (for I = 300 mA) and 60 min of treatment (for I = 30 mA),
therefore much longer for a smaller I value. Then, a rapid decrease took
place, and, finally, it appeared less important luminescence inhibition peaks
at electrolysis times in the 30–180 min region, the electrolysis time values be-
ing generally longer when the current intensity decreased. These time-related
shifts of luminescence inhibition peaks with the current intensity value can
be explained by changes in the rate of formation of the reactive hydroxyl
radicals, resulting from the current intensity changes, and, consequently, in
the degradation kinetics of SMX and its intermediates (Dirany et al., 2011).
Finally this work has shown that the application of the EF process to water
treatment required that particular experimental conditions, namely a high ap-
plied current intensity and electrolysis time values depending on the current
intensity, are fulfilled.
Another original application of the EF process was the determination
of absolute second-order rate constants (k2 ,P ) for the reaction of organic
pollutants with electrogenerated •OH, by applying the competitive kinetic
method (Ammar et al., 2007; Özcan et al., 2008d). This method was based
on the use of a standard competition substrate (S), such as salicylic, benzoic
or p-hydroxybenzoic acids, with a well-known absolute second-order rate
constant (k2 ,S ) for their reaction with •OH. The EF process was then per-
formed with a solution containing both the organic pollutant under study
and the standard competition substrate, and their concentration decay was
simultaneously measured by HPLC analysis to obtain the pseudo-first-order
rate constant of the pollutant (k1,P ) and of the competition substrate (k1,S ).
The value of the pollutant absolute rate constant k2 ,P was then calculated
according to the following equation:
k1,P
k2,P = × k2,S (38)
k1,S
The k2,P values, obtained by this method for several organic pollutants of
different families, were very high, generally ranging from 109 to 1010 M−1 s−1.
These large k2,P values were expected for the hydroxylation reactions of
aromatic compounds, which are characterized by a high reactivity of •OH
radicals versus aromatic organics (Buxton et al., 1988).

5.3 EF-Related Processes: Coupling of EF with Other AOPs

The efficiency of EF process could also be enhanced by its coupling with
other AOPs. Among the coupled processes, the most developed one was the
 Advanced Oxidation Processes in Water/Wastewater Treatment 2617

PEF process, in which the EF reactor was irradiated with an UV lamp. This
PEF method is based on the simultaneous use of electrogenerated H2 O2 in
the presence of Fe2+ (EF process) and of UVA irradiation of the solution to
enhance the mineralization process. The interest of this PEF coupled AOP is
double. First, supplementary •OH radicals were generated by photochemical
reaction of Fe(OH)2+, the predominant species of iron(III) at pH 3 (reaction
(20)), and of H2 O2 (reaction (8)) (Sun and Pignatello, 1993), in addition to
the EF process.
Second, the destruction of Fe(III)-carboxylic acid complexes (reaction
(21)) resulted into ferrous ion regeneration, on the one hand, and in a
significant decay in the TOC content of the system, due to the photodecar-
boxylation of Fe(III) complexes formed with generated carboxylic acids, on
the other hand (Kaichouh et al., 2004; Maezono et al. 2010; De Luna et al.,
2012; Garcia-Segura et al., 2012). For example, reaction (21) shows the pho-
todecarboxylation of Fe(III)-oxalate complexes (Fe(C2 O4 )+, Fe(C2 O4 )2 −, and
Fe(C2 O4 )3 3−) which were formed as ultimate by-products of aromatics.
Consequently, the mineralization degree of the treated solutions was
higher with the PEF process than with the EF process alone. However, this
method also presents a drawback, namely the relatively high electrical cost
of lamps supplying UVA light. Nevertheless, this problem could be solved
by applying the solar photoelectro-Fenton (SPEF) method, in which the
treated solution was directly irradiated with sunlight (wavelengths > 300 nm),
which is a cheap and renewable energy source (Flox et al., 2007; Skoumal
et al., 2009; Khataee et al., 2010; Almeida et al., 2011; Garcia-Segura et al.,
2011a; Isarain-Chavez et al., 2011; Salazar et al., 2012). For example, Skoumal
et al. (2009) performed the degradation of the pharmaceutical ibuprofen in
aqueous medium by the PEF and the SPEF processes at constant current
density and pH 3, in a one-compartment cell with a Pt or a BDD anode,
and an O2 -diffusion cathode. These authors found that the use of sunlight
strongly enhanced the generation of •OH due to a faster regeneration of
Fe(II) in the solution and photodecarboxylation of Fe(III)-carboxylic acid
complexes, which increased the mineralization rate and yield under UVA and
solar irradiation. Results showed that the SPEF process was more potent with
the BDD anode, reaching 92% of mineralization at the end of the treatment.
The operating parameters, such as solution pH, Fe(II) concentration, and
current density, were optimized according to the degradation kinetics and
mineralization efficiency, and the best operating conditions were achieved
by using a Fe(II) concentration of 0.2–0.5 mM, a pH value of 3.0 and a
low current density (6.6 mA cm−2). In these conditions, a mineralization
degree of 86% was obtained for ibuprofen in 3 hrs with an energy cost
as low as 4.3 kWh m−3. In the same study, several aromatic intermediates,
including 1-(1-hydroxyethyl)-4-isobutylbenzene, 4-isobutylacetophenone, 4-
isobutylphenol, and 4-ethylbenzaldehyde, and short-chain carboxylic acids,
such as pyruvic, acetic, formic and oxalic acids, were also identified (Skoumal
2618 M. A. Oturan and J.-J. Aaron

1.2

CO O H
1
O
0.8 Cl

2,4-D,0
/c 0.6
2,4-D,t Cl
c

0.4

0.2

0
0 30 60 90 120 150
time / min

FIGURE 13. Decay of 2,4-D concentration with treatment time during the treatment of 250 mL
of aqueous solutions of 1 mM 2,4D in presence of 0.1 mM Fe3+ as catalyst, at 200 mA, pH 3.0
and room temperature, by using a Pt anode and a carbon-felt cathode. Sonoelectro-Fenton
process with low-frequency ultrasounds (28 kHz), at output power: () 20, () 60, and ()
80 W; () EF process alone. (Source: Oturan et al., 2008)
C Elsevier. Reproduced by permission of Elsevier. Permission to reuse must be obtained from
the rightsholder.

et al., 2009). The characterization of the oxidation reaction intermediates

allowed the authors to propose a plausible reaction mechanistic scheme for
the degradation of ibuprofen by the PEF and solar photoelectron-Fenton
processes. These results confirmed that the PEF process with a gas diffusion
cathode had a higher oxidation power than the EF process applied with the
same cathode. However, it is worthwhile to note that UV irradiation did not
significantly enhance the EF process when a three-dimensional, large-surface
carbon-felt cathode was used.
The sonoelectro-Fenton process is a coupled AOP, in which the treated
solution was simultaneously submitted to the EF process and to ultrasonic
irradiation (Oturan et al., 2008; Li et al., 2010; Babuponnusami and Muthuku-
mar, 2012). Oturan et al. (2008), who invented the term “sonoelectro-Fenton”,
showed the existence of a significant synergetic effect in this AOP between
the sonolysis and the EF process, probably due to the additional effect of
a mass transfer rate enhancement to the electrode by sonification. An un-
divided Pt/carbon felt tank reactor with a ceramic piezoelectric transducer
placed on its base was used to carry out a comparative study on the perfor-
mances of the EF and the sonoelectro-Fenton processes in the degradation
of 2,4-D and 4,6-dinitro-o-cresol herbicides. As can be seen in Figure 13,
the coupled AOP process was enhanced compared to EF alone at low fre-
quencies of 28 kHz with low energy ultrasounds of 20 and 60 W. However,
application of higher energy ultrasounds (80 W) was found to inhibit the
 Advanced Oxidation Processes in Water/Wastewater Treatment 2619

sonoelectro-Fenton process, possibly due to the depletion of O2 needed for

the EF process (Figure 13).
Peroxi-coagulation (or peroxi-electrocoagulation) is another coupled
AOP, mainly developed and applied to wastewater treatment by Brillas et al.
(Brillas and Casado, 2002, 2003a; Paton et al., 2009). This coupled process
involved an EF-like process using a sacrificial iron anode to generate iron(II)
ions according to reaction (39), and a cathode able to yield H2 O2 (reaction
(33)), in order to treat a wastewater sample in an undivided cell.

Fe → Fe2+ + 2e− (39)

Thus Fe2+ ions and H2 O2 were continuously generated electrochemically.

The Fenton’s reaction (32) took place and allowed one to obtain a Fe3+ sat-
urated solution, while the Fe3+ ions in excess precipitated under the form of
an hydrated Fe(III) oxide, (Fe(OH)3 ). Consequently, the pollutants present
in the solution could be removed by oxidation with •OH radicals, produced
during the Fenton’s reaction (1), and by coagulation. It is worthwhile to
stress that the peroxi-coagulation process differed from the classical elec-
trocoagulation with a Fe anode, where soluble organics were not oxidized,
due to the absence of electrogenerated H2 O2 , but only eliminated by coag-
ulation. Moreover, to enhance the process efficiency, the peroxi-coagulation
has been coupled with UV irradiation (photoperoxy-coagulation AOP) (Boye
et al., 2003a, 2003c; Brillas et al., 2003b) for the degradation of chlorophe-
noxy acid herbicides in aqueous media.

6 CONCLUSIONS

Because of the increasing scarcity of the water resources worldwide, and

of the numerous pollutants, generated by agriculture, industries and urban
development, which are disposed of into natural waters, it is very important
and urgent to develop an efficient protection of surface and ground waters.
For this purpose, a good strategy from the technical and economical stand-
point involves treating the wastewaters before their injection in the natural
environment, since only relatively small volumes of toxic and persistent or-
ganic xenobiotics have to be treated, whereas the decontamination of very
large volumes of polluted natural water streams would be very difficult or
even impossible. Following this approach, the development of effective, eco-
logical, economical, and, if possible, easy-to-handle and inexpensive water
treatment technologies are needed.
In this critical review, we have shown that AOPs constitute very effi-
cient technologies for the remediation of waters and wastewaters containing
refractory organic pollutants. Last two decades have been marked by the
2620 M. A. Oturan and J.-J. Aaron

high number and good quality of publications concerning the AOP stud-
ies on mechanisms and applications to water and wastewater treatments,
which demonstrates the great and increasing interest for these methods,
from the theoretical, environmental, and economical points of view. We
have described here the principles, advantages, and drawbacks, as well as
the performances and applications, of the various types of AOPs, including
the chemical AOPs, based on the Fenton reaction and peroxonation, the
photochemical AOPs, namely the photolysis of H2 O2 and O3 , photo-Fenton
process (H2 O2 /Fe2+/UV) and heterogeneous photocatalysis (TiO2 /UV), the
sonochemical AOPs and the electrochemical AOPs, such as AO, EF, and
related processes.
Although the chemical AOPs, including the Fenton’s reagent and perox-
onation, still remain frequently used for a number of applications (treatment
of wastewaters, discoloring effluents of dye industries, and destruction of var-
ious toxic organic compounds such as aromatic compounds, haloalkenes,
and haloalkanes), there has been more recently an increasing number of
works relative to the photochemical, sonochemical, and electrochemical
AOPs, because of the better performances of the later AOPs.
In the case of the photochemical technologies, we can conclude
from our review that the photochemical AOPs are generally simple, clean,
relatively inexpensive, and more efficient than classical, chemical AOPs.
Four main, basic types of photochemical AOPs, namely H2 O2 photolysis
(H2 O2 /UV), O3 photolysis (O3 /UV), photo-Fenton process (H2 O2 /Fe2+/UV),
and heterogeneous photocatalysis (TiO2 /UV), have been applied to de-
grade and/or mineralize organic pollutants in waters. We have pointed out
that, among these photochemical processes, the photocatalytic ones, that
is, photo-Fenton process and heterogeneous photocatalysis, possessed in
most cases a better efficiency than H2 O2 and O3 photolysis. For example,
the photo-Fenton AOP, as compared to the H2 O2 /UV and O3 /UV processes,
was considered by a number of researchers to be by far the best method
for the rapidity of photodecomposition as well as for the mineralization
yield of a variety of pollutants. The solar photo-Fenton AOP, more interest-
ing for reducing the energy consumption, which represents an alternative
to the classical photo-Fenton process, has been also recently utilized for
the removal of various organic compounds present in natural or polluted
waters, degradation of herbicides, and treatment of wastewater effluents
and landfill leachates. Moreover, it is important to stress that heterogeneous
photocatalysis, an AOP mainly based on the use of the semiconductor tita-
nium oxide (TiO2 /UV), has been the object of a tremendous development
in the last decade. Indeed, TiO2 is a material close to practically being
an ideal photocalyst in several important aspects: chemically highly stable,
biologically inert, very easy to produce, inexpensive, active from the pho-
tocatalysis standpoint, and possessing an energy gap comparable to that of
 Advanced Oxidation Processes in Water/Wastewater Treatment 2621

solar photons. For these reasons, numerous environmental and energy ap-
plications have taken place, particularly in the field of water purification,
allowing the oxidation of toxic, inorganic ions, as well as the degradation
and/or mineralization of a number of organic pollutants. Finally, we must
point out that many of the photochemical technologies discussed here, espe-
cially the photocatalytic ones, have the potential to decontaminate wastew-
aters containing a large variety of organic pollutants in a wide range of
experimental conditions. Most of them can be applied to destroy the ini-
tial pollutants, and are frequently able to completely mineralize the treated
solutions.
Concerning the sonochemical AOPs, we can deduce from our literature
search that the combination of ultrasounds with Fenton-type reactions has
resulted into the rapid and recent development of sonochemical methods
for the removal of organic pollutants from waters, and seems able to lead
to a very promising technologic approach for decontamination purposes.
However, most experimental works have been performed until now at the
laboratory scale in artificial systems, and application of sonochemical AOPs
at the industrial level in real-time water (or wastewater) treatment plants
is needed to demonstrate the economic and commercial feasibility of these
sonochemical methods.
The electrochemical AOPs are distinguished from other AOPs by min-
imizing or eliminating the use of chemical reagent. AO process generates
powerful oxidant •OH from oxidation of water on a high O2 overvoltage
anode. The BDD anode emerged as the better anode material by its great
chemical and electrochemical stability, wide electrochemical working range,
and a great oxidation/mineralization power compared to other anodes, since
formed BDD (•OH) is physisorbed on the electrode surface and very reac-
tive. In the case of the EF process, the Fenton reagent (H2 O2 +Fe2+), leading
to the formation of homogeneous •OH via Fenton’s reaction, is electrochemi-
cally generated in situ by a catalytic way. Thus the drawbacks of the classical
Fenton process, such as reagent cost, parasitic reactions, and process sludge
formation, were avoided. As we showed in this review for the atrazine min-
eralization, the EF process became significantly more powerful when the
classical anode (Pt) was replaced by a BDD one, because of the simulta-
neous generation of •OH radicals in bulk solution and on the BDD anode
surface. Both EAOPs (AO and EF) have been successfully applied to the treat-
ment of several types of toxic and/or POPs, like wastewaters contaminated
by pesticides, synthetic dyes, chlorophenols, landfill leachates, and pharma-
ceuticals. Recently, in order to enhance the method performances, several
studies have been performed by coupling EAOPs with other AOPs, leading
to several, new coupled AOPs, such as PEF, SPEF, sonoelectro-Fenton, and
peroxyelectrocoagulation.
2622 M. A. Oturan and J.-J. Aaron

Since they are, in general, economically viable (no or few chemical

reagents), environmental-friendly (use of electrical energy and no process
sludge formation), and technically very efficient (almost complete miner-
alization of pollutants), the EAOPs seem to be very promising water and
wastewater treatment processes and to be considered as technologies of the
future. Now, these technologies have become mature enough to be applied
at the industrial stage

ACKNOWLEDGMENTS

We thank Prof. Dr. Snezhana Efremova Aaron and Dr. Nihal Oturan for their
help in the preparation of this review.

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