Rev Chem Eng 2015; 31(3): 263–302
Archina Buthiyappan, Abdul Raman Abdul Aziz* and Wan Mohd Ashri Wan Daud
Degradation performance and cost implication
of UV-integrated advanced oxidation processes
for wastewater treatments
Abstract: Advanced oxidation processes (AOPs) are commonly used for treating recalcitrant wastewater with varying degree of efficiency, depending on several operating
parameters. In this review, a comparative study among
selected AOPs integrated with ultraviolet (UV) (UV/Fenton, UV/H2O2, UV/O3, UV/TiO2, UV/persulfate, UV/H2O2/O3,
and UV/TiO2/H2O2) was conducted. The cost implication,
changes in kinetics, changes in reaction rates, and effects
of various parameters such as type of contaminants, pH,
catalyst loading concentration of oxidants, and type of
UV light are explained and concluded in this paper. From
this review, it is concluded that UV-integrated AOPs are
efficient for wastewater treatment. However, a few aspects
must be considered including process scale-up, kinetics
of combined processes, reactor configuration, modeling
of a system, and optimization of operating parameters to
enhance the process efficiency.
Keywords: cost evaluation; energy calculation; hydrogen
peroxide; integrated AOPs; recalcitrant wastewater.
DOI 10.1515/revce-2014-0039
Received September 11, 2014; accepted January 29, 2015; previously
published online May 9, 2015
Abbreviations
BOD
COD
CR
DOC
Fe+2
Fe+3
biological oxygen demand
chemical oxygen demand
color removal
dissolved organic compounds
ferrous ion
ferric ion
*Corresponding author: Abdul Raman Abdul Aziz, Faculty of
Engineering, Department of Chemical Engineering, University of
Malaya, 50603 Kuala Lumpur, Malaysia,
e-mail: azizraman@um.edu.my
Archina Buthiyappan and Wan Mohd Ashri Wan Daud: Faculty
of Engineering, Department of Chemical Engineering, University
of Malaya, 50603 Kuala Lumpur, Malaysia
H2O2
HO·
TOC
hydrogen peroxide
hydroxyl radical
total organic carbon
1 Introduction
Advanced oxidation processes (AOPs) depend on the
generation of free radicals such as hydroxyl radical (OH),
hydroperoxyl radical (HO2·), and superoxide radical (O2·-)
(Ayoub et al. 2010, Glaze et al. 1987). Hydroxyl radical
(OH·) plays a central role among these free radicals for
degradation of a variety of recalcitrant pollutants in
AOPs for wastewater treatment (Guittonneau et al. 1990,
Wang and Xu 2012). The importance of hydroxyl radical
in wastewater treatment has been recognized by researchers in the last decades due to its non-selective behavior
which is capable of oxidizing a wide range of hazardous
compounds to carbon dioxide and water or lower-chain
compounds which can then be treated biologically (Guittonneau et al. 1990, Schulte et al. 1995, Arslan et al. 1999,
Trabelsi-Souissi et al. 2011). Furthermore, hydroxyl radical
is a powerful oxidant with oxidation potential of E0 = 2.73 V
and shows faster rates of oxidation as compared to other
conventional oxidants (Zhang et al. 2005, Rosenfeldt et al.
2006).
The most widely discussed AOPs for water and
wastewater treatment are ultraviolet (UV) (Lester et al.
2008, 2012, Avisar et al. 2010), H2O2/UV (Bledzka et al.
2012), ozone/UV (Bustos et al. 2010), ozone/H2O2 (Jung
et al. 2012), ozone/H2O2/UV (Shu 2006), photocatalytical oxidation (Riga et al. 2007), Fenton, and photo-Fenton reaction (Bianco et al. 2011, Duran et al. 2011, Hasan
et al. 2012a,b, Patel et al. 2013). A number of studies have
been conducted using combinations of O3, H2O2 and UV,
ultrasound (US), solar light, and catalysts such as TiO2,
ZnO, Fe2O3, SnO2, ZnS, and CdS (Domínguez et al. 2005,
Riga et al. 2007, Lucas et al. 2010). It has been found that
energy-dissipating components such as US, UV, solar
light, and microwave in combination with oxidants and
catalysts offer better treatment efficiency compared to
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A. Buthiyappan et al.: Degradation and cost of UV IAOP for WT
single systems. AOPs combined with energy-dissipating
components have revealed some advantages such as
increase in the rate of generation of hydroxyl radicals, less
fouling of solid catalyst, higher degradation and mineralization efficiency, complete mineralization, lower reaction
time, better reaction rate, and elimination of mass transfer limitation.
In the literature, many AOP combined technologies
were explored extensively for a variety of pollutants.
Ultrasound in combination with AOPs is one of the most
reviewed combination systems. Recently, Eren (2012) has
reviewed the use of ultrasound with biochemical, electrochemical, ozonation, photolysis, photocatalysis, and
Fenton processes for the degradation of textile dyes and
dye bath. Besides that, Mahamuni and Adewuyi (2010)
conducted a critical review on the cost estimation of AOPs
involving ultrasound for wastewater treatment. It is noted
that Yang et al. (2013) have critically reviewed the application of UV-based AOPs for the treatment of organic
micropollutants in waste and wastewater. This research
was focused mainly on direct UV photolysis, UV/H2O2, and
UV/TiO2 and their efficiency in degrading micropollutants.
Apart from that, there are a number of reviews reported
recently on AOPs such as heterogeneous photocatalytic
oxidation (Gaya and Abdullah 2008) and Fenton oxidation (Garrido-Ramírez et al. 2010, Babuponnusami and
Muthukumar 2013). It is understood that a comparative
review performance of different types of UV-integrated
AOPs varying the pollutants is not reported in the literature. Therefore, this present work is focused on reviewing
the degradation efficiency of different combinations of
UV with AOPS and their cost implication for recalcitrant
wastewater treatment.
Among other energy-dissipating components, UV
integrated with AOPs has received great attention among
researchers working on wastewater treatment. UV oxidation is a destruction process that mineralizes or oxidizes
a variety of organic contaminants found in wastewater by
addition of oxidants and catalysts and irradiation with UV
light. In UV-AOPs system, the contaminants are destructed
by direct oxidation, UV photolysis, and synergistic action
of UV with oxidants and catalysts. As reported in the literature, UV is attaining wide attention for drinking water
treatment, microbial disinfection, and degradation of
organic compounds through direct or indirect photolysis (Alkan et al. 2007, Bin and Sobera-Madej 2012, Lester
et al. 2012). The efficiency of UV light absorption has
been proven to be improvised by the addition of oxidants
or photosensitizing agents (Banat et al. 2005, Jung et al.
2012, Liu et al. 2013). The available literature shows that
radiation has been useful for environmental applications
such as drinking water treatment, wastewater treatment,
decolorization of dyes (Chang et al. 2010), oxidation of
organic compounds, pesticides removal (Chelme-Ayala
et al. 2010), leachate treatment (Hu et al. 2011), pharmaceutical compounds degradation, and palm oil refinery
effluents degradation (Leong and Bashah 2012). On a
separate note, there is growing research focus on application of UV for the degradation of recalcitrant organic pollutants (Azimi et al. 2012, Matafonova and Batoev 2012,
Zoschke et al. 2014).
This paper aims to review the current status of these
effective and still emerging UV-integrated AOPs for
wastewater treatment. In this present work, processes
such as UV/H2O2, UV/O3, photocatalysis, photo-Fenton,
UV/Persulfate, and hybrid methods such as UV/TiO2/
H2O2, UV/TiO2/O3, and other combinations are discussed
in detail. These processes have been given special attention since they have potential to degrade refractory compounds, toxic chemicals, pesticides, and other emerging
pollutants under ambient conditions. Since AOPs are
about generating sufficient hydroxyl radicals to oxidize
the chemical present in the wastewater, it is very crucial
to find a way to accelerate its production for developing
more efficient treatment systems. The free radicals are
effectively produced if the treatment is supplemented
with other energy-dissipating components besides UV
(such as solar and US), combination of more than one
oxidants and catalyst.
This work highlighted the mechanism of the different
processes, operating parameters that contribute to the efficient production of radicals, and the possible combination
of different oxidants, catalysts, and energy-dissipating
components such as microwave and US for better technical feasibility of treatment system in detail. Besides, the
technical and economic feasibility of a process is important to consider as these are the most important aspects of
any treatment system. However, very few studies in the literature have addressed both economic and technical feasibility of the AOPs. Recently, an attempt has been made
by Mahamuni and Adewuyi (2010) to estimate the cost
of ultrasound combined with AOPs for wastewater treatment. Their study showed that the cost of the treatment
system is considerably affected by the type of pollutants,
design of reactor, and the consumption of chemicals. To
date, there is no review reported on the cost estimation of
UV-based AOPs, and in this paper, special attention was
given to the cost calculation of UV AOPs. The possible
components that need to be included when calculating
the capital, maintenance, and operational costs based on
the available research papers are also included. We believe
that this study will help the researcher who is working on
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A. Buthiyappan et al.: Degradation and cost of UV IAOP for WT
the cost evaluation of the UV-AOPs and also to transfer the
laboratory-scale technique to a larger-scale system. This
paper also could help in giving an idea to researchers on
the possible ways of hydroxyl radical production for maximizing the efficiency and minimizing the cost.
2 U
V-based AOPs for treatment
of contaminated water
UV-based AOPs consist of a variety of methods that
show non-selective oxidation of organic compounds.
The main reactive species is the hydroxyl radical, which
degrades many refractory organic pollutants with high
reaction rates. The industrial effluents contain high
chemical and biological oxygen demand; aliphatic and
aromatic hydrocarbons; suspended solids; many different ­macromolecules such as humic acids, sediments, and
chlorinated organic and inorganic salts; high concentrations of bactericides, emulsifiers, and detergents; and
large amount of ammonia, nitrogen, and heavy metals
that are unable to be fully degraded or oxidized by a single
UV system. Therefore, the combination of UV radiation
with other systems is attracting researchers for efficient
removal of recalcitrant organic compounds (Garcia et al.
2007, Lester et al. 2011). This section discusses the most
widely applied UV-based AOPs in sequence: UV/Fenton,
UV/H2O2, UV/TiO2, UV/ozone, UV-persulfate, and various
combinations of UV-AOPs.
2.1 UV/Fenton
Fenton oxidation has been successfully utilized in wastewater treatment for degradation of various hazardous
compounds in the last two decades (Babuponnusami and
Muthukumar 2013). Fenton involves oxidation of ferrous
ion to ferric ion and decomposition of hydrogen peroxide
to hydroxyl radical as shown in Eqs. (1)–(5) (Zhang and
Pagilla 2010).
Fe 2+ + H 2O2 → Fe 3+ + OH - + OH ⋅ (1)
Fe 3+ + H 2O2 → Fe 2+ +⋅O2 H + H + (2)
Fe 2+ + OH ⋅→ Fe 3+ + OH - (3)
Fe 2+ +⋅O 2 H → Fe 3+ +⋅HO2 - (4)
Fe 3+ +⋅O 2 H → Fe 2+ + O2 + H + (5)
Ferric ion can be reduced to ferrous ion in the presence of excess amount of hydrogen peroxide [Eq. (2)].
265
Apart from hydroxyl radical, Fenton reaction also produces ­hydroperoxy radicals (·O2H) which may also be
helpful in the oxidation of organic pollutants. Lately, a
number of studies have been conducted using Fenton
reaction accelerated by UV irradiation. UV-assisted
Fenton process produces more hydroxyl radical compared to conventional Fenton process and is capable of
increasing the degradation rate. Equation (6) represents
the photochemi­cal regeneration of Fe2+ and additional
production of hydroxyl radical by photo-reduction of
aqua-Fe3+ complex (Giroto et al. 2008, De la Cruz et al.
2013).
Fe(OH) 2+ + hv→ Fe 2+ + OH ⋅ (6)
The literature shows that combination of the Fenton
reaction with UV radiation results in better degradation of organic contaminants compared to the typical
Fenton reaction. The investigated organic contaminants
include polymers, pesticides, reactive dyes, EDTA, landfill leachate, sulfonylurea herbicide, oil refinery wastewater, penicillin, ibuprofen, 2-chlorophenol, livestock
wastewater, acetaminophen, malathion pesticide, polyphenols, tea-manufacturing wastewater, palm oil refinery effluent, simulated industrial wastewater, phthalic
anhydride, and naval derusting wastewater (Kusic et al.
2006a,b, Giroto et al. 2008, Zarora et al. 2010, Zhang and
Pagilla 2010). Photo-Fenton process can be carried out
at room temperature and atmospheric pressure (Lu et al.
2012). In addition, iron salt is non-toxic and can be easily
separated in the form of sludge, and H2O2 is an environmentally friendly compound (Liu et al. 2007). In many
studies, photo-Fenton has been reported to be more
efficient for degradation and decolorization of organic
contaminants compared to Fenton treatment (Qiu and
Huang 2010). Besides, Yeber and Cid (2013) recently
reported that UV/Fe system was efficient for removal of
fish oil from fishmeal mill wastewater. The UV/Fe system
successfully achieved good organic matter mineralization while improving effluent biodegradability and
disinfection in their study. The reduction of chemical
oxygen demand (COD) from 5,562 mgO2/l to 1,218 mgO2/l
was observed in their study.
Table 1 summarizes the effectiveness and experimental conditions for removal of various pollutants
from various types of wastewater utilizing photo-Fenton
treatment. Based on the comprehensive studies of photoFenton for wastewater treatment, factors that may affect
degradation efficiency are outlined and discussed in
detail in the following section. In addition, Figure 1 was
also constructed to outline the main operating parameters
and its effects on different type of UV-AOPs.
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UV-C lamp (30 W)
Low-pressure mercury vapor lamp
(20 W, 254 nm)
Azo Dye C.I. Acid Red 14
Mordant Red 73 azo dye
Reactive Dye
Low-pressure mercury UV lamp (30 W,
254 nm)
UV light (15 W UVA lamp (UVP BL-15
365 nm)
Reactive Black B
C.I. Acid Yellow 23
UV-A irradiation (150 W, 360 nm)
Reactive Black 39
A Heraeus UV immersed lamp TN 15/35
with a nominal output of 15 W
Mercury lamp (5 W, UV-C, 254 nm)
Malathion pesticide
C.I. Reactive Yellow 84
Low-pressure mercury vapor lamp of
12 W (254 nm)
Pesticides
Low-pressure mercury lamp (20 W,
254 nm)
Low-pressure mercury lamp (125 W,
UV-C 254 nm)
Low-pressure mercury vapor lamp of
12 W (254 nm)
Pesticide
C.I. Acid Blue 9
UV-lamps
Pollutant
60
30
60
100
180
60
180
45
180
400
120
UV-irradiation
time (min)
Table 1: Summary of studies on the removal of pollutant by (UV/H2O2/Fe).
Elmorsi et al. (2010)
COD: 85
CR: 99
[H2O2]: 20 mm
[Fe2+]: 0.1 mm
[Dye]: 40 mg/l
pH value: 3
UV light intensity: 45.3 W/m2
Modirshahla et al. (2007)
Neamtu et al. (2003)
Daneshvar and Khataee (2006)
CR: 100
CR: 98
COD: 81
TOC: 50
COD: 90
CR: 100
Huang et al. (2008)
Vujevic et al. (2010)
Arslan-Alaton et al. (2009)
CR: 100
COD: 84
TOC: 53
TOC: 98
CR: 100
TOC: 73.8
Zhang and Pagilla (2010)
Malathion removal: > 70
Qiu and Huang (2010)
Abdessalem et al. (2010)
TOC: 93
CR: 98
Abdessalem et al. (2010)
TOC: 90
[Fe2+]: 0.1–1 mm
[H2O2]: 1–100 mm
pH: 3
[Fe3+]: 1 mm
[H2O2]: 50 mm
R = [H2O2]/[Fe3+]: 50 and 100
Malathion:H2O2: 1:100
Molar ratio H2O2:Fe(II): 40:1
pH: 2.5
[Fe3+]: 1.5 mm
[H2O2]: 35 mm
[H2O2]: 647.06 mm
UV irradiation: 45 W
[Fe2+]: 0.72 mm
pH: 3
[Pollutant]: 50 mg/l
[Fe2+]: 0.2 mm
[H2O2]: 10 mm
[Fe2+]: 0.396 mm
[Pollutant]: 5 × 10-2 mm
[H2O2]: 2.5 mm
pH: 3
Temp.: 25°C.
[H2O2]/[Fe2+]: 2.0–3.5
[H2O2]/[Dye]: 7.0
[Dye]: 100 mg/l
[Fe2+]: 0.5 mm
[H2O2]: 2.5 mm
Molar ratio of the Fenton reagent: 1:5
pH: 3
Temp.: 23°C
Molar ratio H2O2/Fe2+: 20:1
pH: 3
References
Treatment efficiency (%)
Experimental conditions
266
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80
160
Low-pressure mercury lamp (8 W, UV-C
254 nm)
Medium-pressure mercury vapor lamp
(30 W, UV-C, 254 nm)
UV-A irradiation (366 nm)
High-pressure mercury lamp (125 W)
High-intensity 254 nm UV grid lamp
Reactive red 241 (RR241)
EDTA
Chlorimuron-ethyl
(sulfonylurea herbicide)
Oil refinery wastewater
UV-lamp with nominal power of 125 W
(wavelength 254 nm)
Xe lamp (290–400 nm wavelength
range)
Medium-pressure Hg lamp of 150 W
(wavelength 254 nm)
Near-UV (black light) fluorescent lamps
(15 W, 352 nm)
Penicillin G
Ibuprofen
2-Chlorophenol
Polyphenols in oolong tea
manufacturing wastewater
Landfill leachate
A Heraeus UV immersed lamp TN 15/35
with a nominal output of 15 W)
C.I. Reactive Red 120
100
45
120
60
240
60
240
30
40
Medium-pressure mercury (15 W,
wavelength 254 nm)
Everdirect supra turquoise
blue FBL-textile dye
UV-irradiation
time (min)
UV-lamps
Pollutant
(Table 1: Continued)
[Dye]: 50 mg/l
[Fe2+]: 0.5 mm, [H2O2]: 2 ml/l
[H2O2]: 5.88×10-2 mm
[Fe2+]: 30 mg l
Temp.: 25°C
pH: 4.0
[EDTA] = 5 mm
[H2O2]: 100 mm
pH: 3.0
[H2O2]: 68.4 mm
[Fe3+]: 0.33 mm
[Fe2+]: 0.14 mm
[H2O2]: 11.77 mm
Molar ratio of the Fenton reagent:
pH: 3
Molar ratio H2O2/Fe2+: 20
[H2O2]: 20 mm
[Fe2+]: 1 mm
Mixing: 120 rpm
pH: 3.5
[H2O2]: 0.32 mm
[Fe2+]: 1.2 mm
pH: 3
pH: 2.5–4.0
[H2O2]: 22 mm
[Fe2+]: 0.45 mm
[H2O2]: 14.71 mm
[Fe2+]: 7.19×10-2 mm
Stock solutions of H2O2 0.95 mm,
FeSO4.7H2O: 0.36 mm
H2O2:Fe2+: 1.63–15.25
pH: 3
H2O2: 5.2
Fe2+: 3.6
Agitation speed: 20 rpm
Molar ratio H2O2/Fe2+: 20:1
pH: 3
Experimental conditions
TOC: 96
COD: 100
COD: 95–97
COD: 60
COD: < 81
TOC: 95.6
CE removal: 94.6
COD: 50
TOC: > 80
CR: 98
COD: 85
TOC: 73
CR: 100
TOC: 18.6
COD: 56.8
TOC: 95
Treatment efficiency (%)
Sabaikai et al. (2014)
Kavitha and Palanivelu (2003)
Mendez-Arriaga et al. (2010)
Saghafinia et al. (2011)
Tony et al. (2012)
Gozzi et al. (2012)
Ghiselli et al. (2004)
Hu et al. (2011)
Patel et al. (2013)
Neamtu et al. (2003)
Liu et al. (2007)
References
A. Buthiyappan et al.: Degradation and cost of UV IAOP for WT
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UV-lamps
UV lamps (40 W, wavelength 254 nm)
UV-C lamps (280–200 nm; 55 W) and
UV-A lamps (400–320 nm; 70 W)
UV-lamp with wavelenghth 254 nm
UV-C emitting irradiation with a
maximum at 254 nm
UV-A emitting irradiation with a
maximum at 365 nm
Low-pressure mercury vapor lamp
(12 W, 254 nm)
Low-pressure mercury UV lamp (UV-A,
365 nm)
Medium-pressure UV lamp (450 W)
Pollutant
Livestock wastewater
Acetaminophen
Palm oil refinery effluent
Simulated industrial
wastewater
Simulated industrial
wastewater
Phthalic anhydride in
aqueous medium
Simulated dyehouse
wastewater
Naval derusting
wastewater
(Table 1: Continued)
60
60
120
60
60
30
120
80
UV-irradiation
time (min)
UV Light Intensity: 5 mW/cm2
Temp.: 20°C
pH: 5
[H2O2]: 0.1 mm
[Fe2+]: 0.01 m
H2O2 flow rate: 50 ml/h,
[Fe2+]: 2 ppm
pH: 2.5
Temp.: 40°C),
H2O2 mass: 2.125 g
H2O2:FeSO4·7H2O: 15:1
Agitation speed: 250 rpm
pH value: 3.88
[Fe2+]: 5.01 mm
[H2O2]: 30 mm
pH: 1.9
[Fe3+]: 8.0 mm
[H2O2]: 30 mm
pH: 3
R = [H2O2]/[Fe3+]: 40
[Fe2+]: 0.1 mm
Temp.: 29±3°C.
[Fe2+]: 0.97 mm
[H2O2]: 30 mm
[Fe2+]: 7.19×10-2 mm
[citric acid]: 0.3 mol/l
[H2O2]: 2400 mm
Experimental conditions
COD: 93
TOC: 61
TOC: 98.7
Kim et al. (2010)
Grcic et al. (2011)
Trabelsi-Souissi et al. (2011)
Dopar et al. (2011)
Dopar et al. (2011)
kobs = 12.94 × 10-2 m-1 s-1
kobs = 9.71 × 10-2 m-1 s-1
Leong and Bashah (2012)
Duran et al. (2011)
Park et al. (2006)
References
COD: > 75
COD: 83
BOD: 94
TOC: 71
COD: 70–79
CR: 70–85
Treatment efficiency (%)
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pHi
[Oxidant]i
[Catalyst]i
• UV/H2O2-mostly
alkaline
• UV/OzoneAcidic (Direct
oxidation,
Ozone), Alkaline
(Indirect
oxidation,OH
radical)
• Fenton based
oxidation - pH3-5
• TiO2/UVAlkaline
• Strongly
dependeny on
amount of H2O2
• Excess amount
causes
scavenging effect
and decrease the
efficiency
• Excessive amout
formed iron
complex
• Reduce the
efficiency by
scavenging the
reaction
• Light
penetration is
reduced by
sludge formation
269
UV light
Intensity
[Pollutant]i
• Concentrated
pollutant need
more dosage of
oxidant as well as
catalyst.
• Tubidity of
pollutant,
decrrease the
amount of UV
light that pass
through the
system
• Proper choice of
UV source
• Turbidity effects
the light
penetration
• Optimize the
design of
photoeractor for
efficient
distribution of
light
Figure 1: Operating parameters and their effects on treatment efficiency of UV-AOPs.
2.1.1 Factors affecting wastewater treatment using
photo-Fenton process
2.1.1.1 Operating pH
The hydroxyl radical generation by photo-Fenton processes is strongly dependent on the initial pH of the solution since pH value has a significant effect on the oxidation
potential of OH radicals. This is due to the inverse relation
between the oxidation potential and pH value (Eo = 2.8 V
and E14 = 1.95 V) (Lide 2004). At acidic conditions, efficiency of photo-Fenton process is improved regardless
of types of pollutants to be degraded (Philippopoulos
and Poulopoulos 2003). This is because acidic conditions
favor the formation of OH radical, whereas H2O2 is decomposed into O2 at higher pH values, and it loses its oxidation ability. However, extreme acidic conditions are also
not suitable for efficient OH radical generation. Hydrogen
peroxide gets solvated and becomes more stable at pH
below 2, reducing its reactivity with ferrous ion to generate
hydroxyl radical (Duran et al. 2011). Axonium ion (H3O2)+
is formed at higher concentration of H+ making hydrogen
peroxide more stable and preventing it from reacting with
ferrous ion (Feng et al. 2006). Therefore, pH value needs
to be optimized for efficient production of OH radical to
improve the overall degradation efficiency.
Besides OH radical, availability of ferrous iron in
reaction medium is also influenced by pH value. Different
forms of iron species in relation to pH values are summarized in Table 2. It is clear from the table that at very low
pH, iron is present in the form of ferric iron complex which
acts slowly with hydrogen peroxide to form OH radical
(Lucas and Peres 2006). Most of the ferrous iron precipitates as Fe (OH)3 and forms amorphous oxyhydroxides
(Fe2O3·nH2O) at higher pH (Ghiselli et al. 2004). Presence of
iron precipitates does not only decrease the light absorption but also increase the cost of post-treatment process.
Most of the studies concerning Fenton and photo-­
Fenton processes have suggested pH = 3 as the optimum
pH value. For instance, Lucas and Peres (2006) studied the
decolorization of azo dye Reactive Black 5 by photo-Fenton oxidation within pH range of 1–3. The color removal
efficiencies of 32.6%, 61%, and 98.6% were obtained at pH
1, 2, and 3, respectively. The result showed that maximum
color removal was obtained at pH 3. Lower degradation at
pH 1 and 2 was caused by the hydrogen scavenging effect
due to excessive H+ as mentioned earlier. In contrast,
Park et al. (2006) have discovered pH = 5 as the optimum
pH value for degradation of livestock wastewater through
photo-Fenton process. The authors commented that
hydrogen peroxide was the most stable in the range of
pH 3–4, but the decomposition rate increased rapidly at
pH 5 and decreased above pH = 5. Therefore, the optimum
pH was fixed at 5 for this study. In addition, Dopar et al.
(2011) highlighted the effect of UV type on pH for treatment of simulated industrial wastewater. It was observed
that optimum pH value varied with UV type used. The
highest degradation rate was obtained at pH of 3.88
([Fe2+] = 5.01 mm and [H2O2] = 30 mm) for UV-C irradiation,
while pH of 1.9 ([Fe2+] = 8.39 mm and [H2O2] = 30 mm) was
Table 2: Iron species generated in photo-Fenton as a function to pH
value (Neamtu et al. 2003).
Fe3+ species
Fe(H2O)63+
Fe(OH)(H2O)52+
Fe(OH)2(H2O)4+
pH range
1–2
2–3
3–4
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A. Buthiyappan et al.: Degradation and cost of UV IAOP for WT
observed as optimum pH for UV-A light. From the literature, photo-Fenton also shows efficient results in acidic
conditions like Fenton process, but it has to be optimized for each type of UV system, as evident from results
obtained by Dopar et al. (2011).
2.1.1.2 Ferrous ion concentration
The amount of ferrous ions is one of the main factors
influencing the process due to its action as a catalyst to
decompose hydrogen peroxide to generate OH radicals.
According to previous studies, the minimum ferrous ion
concentration required for the reaction is between 3 and
15 mg/l. As reported by Vujevic et al. (2010) and Tony et al.
(2012), degradation rate increased with an increase in
concentration of ferrous ion till a certain concentration,
and it became inefficient above that value.
Increasing ferrous ion above the optimal value has
adverse impacts on reaction, where it may act as hydroxyl
radical scavenger according to Eq. 7 (Vujevic et al. 2010).
Excessive Fe2+ can result in a great self-consumption of
free radicals, which can inhibit the oxidation reaction as
presented in reaction Eqs. (1) and (2).
Besides, enormous increase of ferrous ion leads to
formation of brown turbidity, which hindered the absorption of UV light and made the process efficiency drop
(Herrera-Melián et al. 2000). Furthermore, the sludge
formed requires post-treatment, and this is not economically viable (Tony et al. 2012). So the concentration
of ferrous ion should be as low as possible to minimize
the formation of sludge. This is because sludge removal
requires post-treatment, and this is not economically
viable (Tony et al. 2012).
Fe 2+ + OH ⋅→ HO- + Fe 3+ (7)
Vujevic et al. (2010) conducted a study on decolorization
and mineralization of reactive dye by UV/Fenton process.
The effects of Fe2+ concentration on dye degradation efficiency were investigated within the range from 0.05 to
1.0 mm in this study. The results indicated that mineralization increased from 35% to 73.8% with an increase in
the initial Fe2+ concentration from 0.05 to 0.5 mm. Therefore, it can be concluded from the results that degradation
rate decreased when iron concentration is above 0.5mm.
The degradation of terephthalic acid (TPA), isophthalic
acid (IPA), and benzoic acid (BA) from terephthalic acid
wastewater by AOPs was evaluated by Thiruvenkatachari
et al. (2006). The authors have reported that enhancement in the removal efficiency was observed for all three
target organic species in the photo-Fenton process compared to UV/H2O2 system due to more hydroxyl radical
in UV-assisted Fenton. Besides, it was found that the
removal efficiency increased with increase in Fe concentration. The rate of organic destruction is faster in
the early stage of the reaction than in the later stage. It
is because ferrous ion catalyses H2O2 quickly in the first
stage of reaction to form hydroxyl radical, so more degradation of organic compounds occurs in the early stage of
the reaction. In this study, selectivity of hydroxyl radicals
for certain specific organic materials was observed. Complete removal of BA was observed within a short period
of reaction time when only about 60% of TPA and IPA
were degraded.
The amount of ferrous ions is one of the main parameters that influence the efficiency of the photo-Fenton
process. On the other hand, ferrous ion concentration
depends on the type of organic contaminants present in
the wastewater and therefore should be optimized for
each particular treatment system.
2.1.1.3 Hydrogen peroxide concentration
The concentration of hydrogen peroxide contributes to the
degradation efficiency of the targeted pollutants in photoFenton process. Hydrogen peroxide can generate powerful
hydroxyl radical by acting as an electron acceptor as in
Eq. (8) (Alkan et al. 2007, Elmorsi et al. 2010). The absorbance of UV light by hydrogen peroxide can generate two
hydroxyl radicals as shown in Eq. (9)
H 2O2 +e- →OH ⋅+OH - (8)
H 2O2 +hv→OH ⋅+OH ⋅ (9)
Based on the observed result, degradation efficiency
increased with the amount of hydrogen peroxide. However,
the concentration of hydrogen peroxide should be selected
carefully to avoid excessive hydrogen peroxide as it can
decrease the degradation efficiency or does not improve
the degradation. This might be due to auto-decomposition
of hydrogen peroxide to oxygen and water together with
recombination of hydroxyl radical, which results in scavenging effect as shown in (10) and (11).
2H 2O2 → 2H 2O+O2 (10)
H 2O2 + OH ⋅→ H 2O+HO2 ⋅ (11)
HO2 ⋅+OH ⋅→ H 2O + O2 (12)
OH ⋅+OH ⋅→ H 2O2 (13)
Besides, dosage of hydrogen peroxide highly depends
on the type of organic contaminants, so it is very important to optimize the value based on the treatment. Papic
et al. (2009) evaluated the degradation of three different
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reactive dyes using homogeneous photo-Fenton process.
The optimum H2O2 concentration was found to be 20 mm
for C.I. Reactive Yellow 3 (RY3), 2.5 mm for C.I. Reactive
Blue 2 (RB2), and 50 mm for C.I. Reactive Violet 2 (RV2),
which is in accordance with the optimum molar ratio of
Fe2+/H2O2.
Duran et al. (2011) evaluated the photo-Fenton mineralization of synthetic municipal wastewater effluent containing acetaminophen in a pilot plant. The
results showed that the degradation rate continuously
increased with the peroxide flow rate, as more ·HO
radicals were being generated. Further analysis was
conducted to understand the effects of hydrogen peroxide by analyzing the remaining and consumed H2O2
in the solution. The increase in H2O2 flow rates was not
proportional to the increase in degradation rate. This
might be due to formation of oxonium ion (H3O2+) and
intermediate product through H2O2reaction with organic
pollutants. H2O2 contributes to hydroxyl radical scavenging capacity at higher H2O2 dosages. Therefore, optimal
H2O2 should be determined to achieve the most efficient
degradation.
2.1.1.4 Initial concentration of pollutant
Initial pollutant concentration plays a significant role in
efficiency of photo-Fenton process. It has been observed
that degradation potential of UV decreases with increase
in pollutant concentration (Zhang and Pagilla 2010). This
is because light penetration decreases with increase in turbidity. It does not only result in decrease in photo-degradation rates of targeted contaminant but also significantly
reduces the photo-reduction of ferric ions (reaction 6) and
H2O2 to OH radicals (reaction 9).
Tokumura et al. (2013) evaluated efficiency of photoFenton process for simultaneous color wastewater treatment. The initial Orange II concentration varied from
22 to 78 ppm to determine the effects of initial Orange II
concentration. All the other operating parameters were
kept constant with the initial hydrogen peroxide concentration of 297 ppm, initial total iron ion concentration of
10 ppm, a solution pH of 2.5, and a total UV-A light intensity of 94.7 W/m2. It was reported that k decreased with
increasing initial Orange II concentration from 22 ppm
(k = 0.445 min-1) to 78 ppm (k = 0.334 min-1).
From the above studies, it can be concluded that
photo-Fenton process becomes inefficient at higher concentration of pollutant. In this situation, more hydroxyl
radicals are required for efficient degradation which can
only be possible through utilizing high concentrations of
iron salt and hydrogen peroxide.
271
2.1.1.5 Operating temperature
Based on the literature review, there is a limited number
of studies that have highlighted effects of temperature on
degradation efficiency. The generation of OH· and HO2·
can be promoted by increasing reaction temperature,
but the generation of OH radical is not favorable at high
temperature. It can be supported by the study conducted
by Zhan et al. (2013). This study evaluated the effects of
solution temperature ranging from 30°C to 70°C on the
oxidation of mercury. The result illustrated that oxidation efficiency increased when the solution temperature
increased from 30°C to 40°C, but there was a sharp decline
following further increase of the solution temperature to
70°C.
Torrades and García-Montaño (2014) also investigated
the optimization of parameters for degradation of real
dye wastewater by photo-Fenton reactions. The degradation studies of Tartrazine were conducted at different
temperatures in the range of 298–333 K under optimum
experimental conditions with pH = 3, [H2O2] = 73.5 mm, and
[Fe(II)] = 1.79 mm. The rate constant increased with temperature till the temperature reached 323 K. The rate constant decreased above this temperature. This was a result
of decomposition of hydrogen peroxide into water and
oxygen at higher temperature.
Thus, based on the previous studies, increasing temperature can be considered a way of intensification of the
Fenton process. But the temperature and amount of H2O2
should be adjusted effectively to achieve a high mineralization. More studies need to be conducted to see the
effects of higher temperature on the amount of ferrous
ion, mechanism of reaction, and intermediate generation.
2.2 UV/H2O2
Photolysis of hydrogen peroxide (H2O2) using UV light
remains the most often investigated advanced oxidation
method for water and wastewater treatment. UV/H2O2 is
still a promising technique because no sludge is produced
at the end of reaction with easier handling and shorter
reaction time to degrade contaminants compared to other
AOPs (Aleboyeh et al. 2005, Autin et al. 2012). Besides,
UV/H2O2 system is the most appropriate AOP technology
for removing toxic organics since hydrogen peroxide is
cheaper and requires fewer safety precautions.
Oxidation in UV/H2O2 system may occur via one of
three general pathways such as hydrogen abstraction,
electron transfer, and radical addition. The reaction
between UV light and hydrogen peroxide generates powerful hydroxyl radical (·OH) as shown in reaction (14)
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A. Buthiyappan et al.: Degradation and cost of UV IAOP for WT
(Xu et al. 2011, Zuorro and Lavecchia 2013). Radiation
with a wavelength < 400 nm is suitable to photolyze H2O2
molecule into hydroxyl radicals with a yield of two ·OH
radicals per quantum of radiation absorbed [Eq. (14)].
The other reactions also occur depending on the process
condition [Eqs. (15)–(17)]. The photolysis reaction of H2O2
depends on the conditions of the system including temperature, initial concentration of H2O2, pollutant, and pH
(Malik 2004, Banat et al. 2005, Shu et al. 2006b, Yonar
et al. 2006, Li et al. 2011a).
H 2O2 + hv→ 2⋅OH (14)
H 2O2 + HO⋅→ H 2O+HO2 ⋅ (15)
HO⋅+ HO2 - → HO2 ⋅+OH - (16)
H 2O2 + HO2 ⋅→ HO⋅+ H 2O2 + O2 (17)
According to previous studies, degradation rate of
UV/H2O2 strongly depends on pH (Shu and Chang 2005a,
Muruganandham and Swaminathan 2006a). At high pH
value, hydrogen peroxide deprotonates and generates
H2O2/HO2- equilibrium (Aleboyeh et al. 2005). As shown
in reaction (18), non-dissociated molecule of H2O2 reacts
with HO2- species and is broken down to oxygen and water
instead of hydroxyl radical (Schrank et al. 2007). Decomposition of H2O2 decreases the amount of hydroxyl radical
available and eventually decreases degradation rate.
Besides, hydrogen peroxide shows a very high self-decomposition rate at higher pH (Shu and Chang 2005b, Schrank
et al. 2007). As a conclusion, raising pH results in reducing dye degradation rate since H2O2 dissociates into water
and oxygen rather than hydroxyl radicals in alkaline condition. So it is necessary to determine the optimum value
of pH for each treatment system investigated.
HO2 - + H 2O2 → H 2O + O2 + OH - (18)
Rauf et al. (2008) also studied the photolytic decolorization
of Rose Bengal by UV/H2O2 and observed that pH played
an important role in degradation. The study showed that
the maximum dye decolorization was 90% at pH 6 for dye
concentration of 0.005 mm and H2O2 of 0.042 mm. The rate
of dye decolorization was low in highly acidic solutions.
The maximum dye decolorization was achieved at pH 6.6,
and the decoloration reduced at alkaline pH. The authors
attributed the enhanced dye decolorization in less acidic
environment to formation of OH radical when peroxide
anions were irradiated by UV light (15). The result was also
supported by the study conducted by Jiraroj et al. (2006).
The authors observed a rapid degradation of PbEDTA in
acidic solutions, while lead precipitation was achieved at
pH higher than 6. Slower degradation was observed at high
pH conditions because a small number of H2O2 is depronotaed to yield HO2- anion. The generated anion acts as a
hydroxyl radical scavenger since this HO2- anion absorbs
UV and generates ·OH with a higher molar absorption
coefficient compared to that of H2O2. Besides, the reaction
between ·OH and HO2- is faster than that of H2O2 and HO2-.
Chelme-Ayala et al. (2010) also investigated the degradation of pesticides in natural water by UV irradiated H2O2
system. At an initial H2O2 concentration of 8.8 × 10-4 m, the
degradation of pesticides at different pH was investigated.
The authors observed that pH had little effects on degradation efficiency of pesticides. In addition to the above
mentioned investigation, Schrank et al. (2007) showed
the important role of pH for efficient degradation and
necessity of pH optimization for each UV/H2O2 system as
it varies with pollutant types.
Initial concentration of H2O2 may either enhance
or inhibit the photoreaction rate of a UV/H2O2 system. A
study conducted by AlHamedi et al. (2009) discovered
that the rate of dye decoloration was directly proportional
to H2O2 concentration till a certain concentration. Above
that value, the increase in dye decoloration was not linear.
This is because the solution undergoes self-quenching of
·OH radicals with excess amount of H2O2 at higher concentration and produces HO2· radicals as shown in Eq. (19).
H 2O2 +⋅OH → H 2O + HO2 ⋅ (19)
In addition, Elmorsi et al. (2010) also studied the effects
of initial concentration of H2O2 (2.5 × 10-5, 2.5 × 10-3, 2.5 × 10-1,
and 5 × 10-1 m) on photo-degradation efficiency of an H2O2/
UV system using fixed concentration of Mordant red 73 azo
dye solution (5 × 10-2 mm) at pH 3.0 and 25±2°C. The highest
degradation rate of the dye was observed with H2O2 concentration of 2.5 × 10-3 m (rate constant of 0.0863 min-1). Further
increase of concentration of H2O2 to 5 × 10-1 m inhibited the
reaction, as confirmed by a decrease in the rate constant
to 0.0103 min-1. Therefore, the optimum concentration of
H2O2 in the reaction course must be reached to maximize
the degradation rate. Muruganandham and Swaminathan
(2006a) also investigated the effect of addition of H2O2
(5–25 mmol/l) on decolorization of Reactive Yellow 14. It
was found that addition of 5–20 mmol/l of H2O2 increased
the decolorization from 40.1 to 60.1%. However, a further
increase of H2O2 to 25 mmol/l decreased the decolorization from 60.1 to 59.2% within 60 min. Enhanced decolorization was observed by adding H2O2 up to 20 mmol/l
due to increase of hydroxyl radical. However, at high
H2O2 concentration, hydroxyl radicals act as a quencher
and causes decrease in treatment efficiency. In addition,
Lester et al. (2010) also studied the photodegradation of
the antibiotic sulphamethoxazole (SMX) in water using
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a medium-pressure UV lamp combined with H2O2. They
observed, at higher concentration of hydrogen peroxide
(50–150 mg/l), that the efficiency of pollutant degradation
did not show any significant improvement in comparison
with 50 mg/l of H2O2. The authors concluded that concentrated H2O2 may decrease the availability of light for direct
photolysis of pollutant, and scavenging of the hydroxyl
radicals by H2O2 also causes the efficiency to drop at
higher concentrated H2O2. These observations are also in
agreement with the result of other studies (Malik 2004,
Shu and Chang 2005a, Riga et al. 2007, Li et al. 2008).
UV radiation is another prominent parameter which
has significant effects on degradation of pollutants in UV/
H2O2 system besides initial concentration of H2O2 and pH
(Muruganandham and Swaminathan 2006a). Haji et al.
(2011) investigated degradation of methyl orange (MO) dye
using AOP by exposing MO aqueous solution to UV irradiation after addition of hydrogen peroxide (H2O2). AOP
process was carried out by studying the influence of dye
concentration, H2O2 dosage, UV irradiation power, and the
area of the solution exposed to the UV source. The effect of
irradiation power was investigated by changing the power
from 0 to 30 mW. The result showed that linear increase
in dye degradation rate was observed with increase in UV
irradiation power. Increase in UV radiation power caused
a rise in the formation of hydroxyl radical from hydrogen
peroxide and increased the degradation rate.
On top of that, the type of UV lamp also affects pollutant degradation in UV-based processes. Linden et al.
(2007) studied the effectiveness of UV/H2O2 processes
using low and medium-pressure lamps to degrade endocrine-disrupting chemicals (EDCs). The authors observed
that the emission spectrum of a medium-pressure mercury
UV lamp covered most of the major absorbance range of
contaminants investigated. Therefore, the authors concluded that a medium-pressure mercury UV lamp was
more effective than a low-pressure mercury lamp. In this
study, 17-alpha-estradiol (E2) and 17-beta-ethinyl-estradiol, suspected EDCs, were degraded between 80 and
99.3% at H2O2 concentration of 15 ppm and a UV dose
of 1000 mJ/cm2 using either low- or medium-pressure
lamps. On the other hand, Rosario-Ortiz et al. (2010) used
a custom-made low-pressure (LP) UV collimated beam
system in UV/H2O2 process to degrade six pharmaceuticals (meprobamate, dilantin, carbamazepine, primidone,
atenolol, and trimethoprim). They found that the removal
of the contaminants correlated with the reduction in UV
absorbance. Higher dose of UV radiation was needed to
promote the generation of hydroxyl radical to overcome
the scavenging effects and increase the oxidation efficiency of pollutants. They also observed that the choice
273
of UV lamp depends on the type of pollutant and the
­system’s condition.
In addition, a number of researchers have also investigated the efficiency of UV/H2O2 in pilot-scale systems. De
la Cruz et al. (2013) investigated the potential use of UV
alone, UV/H2O2, and UV/Fenton to treat micro-pollutants
in a continuous effluent from a municipal wastewater
treatment plant (MWTP) at pilot scale. They observed degradation > 80% with UV irradiation and H2O2 within very
short reaction times. The degradation threshold was also
achieved within the retention time of 10–67 s, and the
maximum amount of H2O2 needed was only 50 mg/l.
Ghafoori et al. (2014) proposed a methodology for
the photo-reactor scale-up, which included all reaction
mechanisms for binary degradation of aqueous polyvinyl
alcohol (PVA) by UV/H2O2 process. They determined the
intrinsic kinetic parameters using the experimental data
obtained in a laboratory-scale batch recirculating system.
The performance of a pilot-scale continuous-flow photoreactor was predicted using the kinetic parameters influencing the photochemical reaction of PVA in a batch-scale
reactor without any parameter adjustment. Such kinetic
parameters were used because they are independent of
the photoreactor geometry and operation. Therefore, they
can be used to predict the performance of the pilot-scale
photoreactor with different geometry and mode of operation. The authors highlighted three important conditions
to use this developed method including a valid kinetic
scheme, detailed mechanistic kinetic model to identify
the intrinsic reaction kinetic expression, and an accurate
mathematical model. In this study, computational fluid
dynamics (CFD) was used to simulate the results of the
proposed model. A good agreement between the CFD simulation result and the experimental data was observed,
and thus the reliability of this scaled-up methodology was
confirmed.
A pilot-scale plug-flow photo-reactor (UV/H2O2) was
used to treat azo dye acid orange 10 by Shu and Chang
(2005c). The key parameters in this experiment, including flow rate, hydrogen peroxide concentration, UV input
power, pH, and initial dye concentrations, were investigated. Ultimately, the obtained result was compared with
the ones from a batch reactor, and it was observed that the
degradation rate of dye in a pilot-scale plug-flow photoreactor (UV/H2O2) was 233 times higher than that of batch
reactor with the same UV light source. Besides, the time
needed to achieve 99.9% of color removal of 20 mg/l acid
orange 10 dye was only 26.9 min, which was much shorter
than that of batch reactor which required 186.4 min to
obtain < 1.0% of residual color. Ninety percent decolorization was reached at the flow rate of 1.63 l min-1 with lower
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hydrogen peroxide concentration of 0.233 mm. The result
showed that light penetration was the most important
factor for enhancement of degradation rate in UV/H2O2 and
any UV-based processes. The pilot-scale system showed a
better removal efficiency compared to batch-scale reactor.
It is due to the larger gap that the batch reactor resulted
in ineffective decolorization. It is therefore suggested that
it is important to reduce the gap between the reactor wall
and UV lamp, especially in large-scale systems for better
degradation efficiency. This is because there is less light
penetration when the gap is larger, causing less hydroxyl
radical to be produced and decreasing the efficiency of the
process.
More applications of UV/H2O2 for wastewater treatment is presented in Table 3. As discussed earlier, the
effectiveness of UV/H2O2 process strongly depends on
factors such as UV radiation, retention time, initial pH
of the solution, initial concentration of hydrogen peroxide, initial concentration of pollutant, and type of pollutant. It is noted from the literature that all parameters
have significant impacts on the efficiency of the UV/
H2O2 process in batch- or large-scale system. UV/H2O2
has been successfully used to degrade different types of
pollutants including synthetic dyes, landfill leachate,
pharmaceutical, emerging pollutants [diethyl phthalate
(DEP), chlorophenols, and clofibric acid], and domestic
wastewater. The results showed that complete decolorization was achieved by using UV/H2O2 with the COD of
the sample being successfully reduced from 40 to 90%.
Besides, pH 3–5.5 has been found to be suitable for synthetic dye’s degradation, and the value varies with types
of dyes. On the other hand, degradation of pharmaceutical, methyl tert-butyl ether (MTBE), and chorophenolcontaining wastewater has been found to be efficient at
alkaline conditions. The differences in pollutant removal
efficiency in acidic and alkaline solution may be attributed to formation of intermediates or presence of certain
species in the reaction solution. The initial concentration
of H2O2 also varies with the concentration of pollutants
and radiations applied in the system. Furthermore, it is
noted from Table 3 that low-pressure lamp has been used
most frequently in UV/H2O2 systems for different types of
pollutants. Therefore, optimization of all the above mentioned parameters is deemed important to improve the
efficiency of a treatment system.
Based on the literature, it is clear that UV/H2O2 treatment is one of the promising technologies for wastewater
treatment. However, wastewater with high COD and total
organic carbon (TOC) value may not be easily treated with
this system. There may also be a need for pre-treatment to
enhance the degradation as suggested by Li et al. (2008).
Besides, UV-H2O2 is relatively expensive compared to other
conventional methods. The process should therefore be
optimized to become an industrially applicable and costeffective technology for organic pollutant degradation.
2.3 UV/TiO2
For years, the potential of photocatalysis for organic degradation using TiO2 has been extensively studied. Photocatalytic oxidation takes place when titanium (IV) oxide
is irradiated by UV light (Autin et al. 2012) with energy
higher or equivalent to band gap. As a result, electrons in
the valence band excite to the conduction band (LydakisSimantiris et al. 2010). This mechanism results in the formation of a positive hole (H+) in the valence band and an
electron (e-) in the conduction band (Shavisi et al. 2014).
Species like O2- and OH radicals are produced as a result
of O2 and water interaction with the photo-catalyst which
helps in degradation of organic pollutants. The principal
mechanisms of photocatalytic oxidation are shown in Eqs.
(20)–(23).
TiO2 + hv-→ H + + e- (20)
e- + O2 →O2 - (21)
H + + H 2O → HO·+ H + (22)
HO + organic →CO2 (23)
The degradation mechanism of TiO2 combined with
UV treatment that involves indirect photolysis and heterogeneous photocatalysis makes the decomposition of
pollutants efficient. UV light with short wavelengths and
high energy makes the two reaction pathways possible
through OH radical oxidation and electric holes reaction.
In the past decade, TiO2 and ZnO have been widely used by
numerous authors for degradation of organic matter, and
TiO2 has been recognized as a benchmark semiconductor
for effective degradation of organic pollutants (Daneshvar
et al. 2004, Peternel et al. 2007). TiO2/UV methods have
gained considerable acceptance due to their ability to
completely degrade a wide range of pollutants. Organic
matter that can be successfully treated using TiO2/UV
includes synthetic dyes, polyaromatic hydrocarbons
(PAHs), volatile halogenated organic solvents, chlorophenols, biphenyls, 2-chloroaniline, 2,4-dichlorophenol,
amoxicillin, DEP, lignin wastewater, municipal wastewaters, and pesticides (Tang et al. 1997, Daneshvar et al.
2003, Park et al. 2003, Chang et al. 2004, El Hajjouji et al.
2008, Lin et al. 2012, Mohammadi and Sabbaghi 2014). It
should be noted that initial concentration of pollutant,
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Mercury lamp (45 W, 254.7 nm)
Medium-pressure mercury vapor lamp
(30 W, UV-C, 254 nm)
Monochromatic lamps (254 nm) with
nominal power of 15 W
Olive mill wastewater
Landfill leachate
UV lamps (365 nm and 254 nm)
UV lamps (365 nm and 254 nm)
Low-pressure mercury lamp (254 nm)
Low-pressure mercury lamp (15 W,
254 nm)
Clofibric acid
Diclofenac
Humic surface waters
MTBE removal
3
120
Low-pressure, mercury vapor
sterilization lamp (40 W)
Low-pressure mercury lamp (8 W)
Low-pressure mercury lamps (254 nm)
Low-pressure mercury lamp (254 nm)
Low-pressure UV lamp (10 W, 254 nm)
Commercial naphthalene
sulphonate
Melatonin
Herbicides simazine
Pathogen re-growth in
UASB effluent
Cyclohexanoic acid
40
60
180
Low-pressure mercury vapor lamps
(254 nm)
60
30
5
60
60
70
300
180
60
60
UV-irradiation
time (min)
4-Chloronitrobenzene
(CNB)
Chlorophenols
Seawater
Low-pressure mercury vapor UV lamp
(254 nm, 15 W)
Low-pressure mercury lamp (254 nm,
4.9 W)
Low-pressure mercury arc lamp
(254 nm)
Pharmaceutical wastewater
Domestic wastewater
UV-lamps
Pollutant
Table 3: Summary of studies on the removal of pollutant by (UV/H2O2).
[H2O2]: 1.765 mm
Irradiation: 5 mW/cm2
Irradiance: 0.11 mW/cm2
[H2O2]: 5 mm
[CHA]: 10 mg/l
Molar ratio of H2O2/CHA: 26
[Bacteria]: 106–107 CFU m/l
Temp.: 23°C
[CPs]: 0.1 mm and 0.4 mm
[H2O2]: 5–40 mm
pH: 9.5
[Clofibric acid]: 0.0467 mm
[H2O2]: 1 mm
[Diclofenac]: 0.0314 mm
[H2O2]: 1 mm
[H2O2]: 3.68 × 10-3 mm &
8.82 × 10-2 mm
UV light dose: 68–681 mWs/cm2
[MTBE]: 0.01–0.06 mm
Molar ratio H2O2/MTBE: 30 and 100
pH: 6–7
[CNB] = 2.5 × 10-6 mol/l
Flow rate: 60 l/h
pH: 7.5
[H2O2]: 0.2 mm
pH value: 5.9–6.0
[H2O2]: 60 mm
[Melatonin]: 20 mg/l
[H2O2]: 10 mm
pH: 4
[H2O2]: 3.53 mm
[Wastewater]: 100 mg/l
Temp: 27–37°C
[H2O2]: 20 mmol/dm3
[H2O2]: 20 mm
Temp.: 25°C
pH: 3
[H2O2]: 117.65 mm
Temp.: 22°C
[H2O2]: 1.471 mm
Experimental conditions
Yonar et al. (2006)
COD: > 95
Alkan et al. (2007)
Bacteria reduction: 7 log
Yasar et al. (2007)
Afzal et al. (2012)
k: 2.2 × 10-2±0.003
Li et al. (2011a)
Xu et al. (2009)
Tureli et al. (2010)
Guittonneau et al. (1990)
Kinetic
constant: maximum value
0.1701 min-1
Pathogens Removal: > 99
COD: 48
TOC: 27
Melatonin
Degradation: > 85
CNB removal: 55
Alnaizy and Ibrahim (2009)
Kim et al. (2013)
TOC: 60
Degradation: 100
Kim et al. (2013)
Trapido et al. (1997)
Degradation: 100
TOC: 80
Rubio et al. (2013)
Disinfection = 99.9
Bin and Sobera-Madej (2012)
Hu et al. (2011)
COD: 48.2
COD: 7.6 × 10-4 s-1
Drouiche et al. (2004)
References
COD: 94
Treatment efficiency (%)
A. Buthiyappan et al.: Degradation and cost of UV IAOP for WT
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Low-pressure Hg arc-UV lamp (254 nm)
Low-pressure mercury lamp ( 10 W,
254 nm)
Low-pressure mercury lamp (6 W,
254 nm)
Medium and low-pressure mercury
lamp (15 W, 1000 V, 220 V)
Low-pressure mercury lamp (254 nm)
Amoxicillin
Clofibric acid (CA)
Chloramphenicol
Reactive black 5
Low-pressure mercury arc UV lamp
(253.7 nm, 14 W)
Low-pressure mercury lamp (15 W)
UV lamp (UV power output of 6 W at
254 nm)
R-52 Mineralight® lamp (254 nm)
Medium-pressure mercury lamp
(125 W)
Low-pressure mercury arc UV lamp
(253.7 nm, 14 W)
Acid Blue 74
Rhodamine B
Methyl orange
Textile wastewater
Textile dyeing wastewater
Low-pressure mercury arc UV lamps
(16 W, 253.7 nm)
C.I. Direct Blue 199
C.I. Acid Blue 113
Medium-pressure Hg vapor lamp
(254 nm)
Textile-dyeing wastewater
C.I. Acid Blue 25
UV-lamps
Pollutant
(Table 3: Continued)
180
180
3
30
7
20
30
10
40
65
50
15
80
UV-irradiation
time (min)
[H2O2]: 10 mm
Temp.: 22°C
[Dye]: 10.0 mg/l
[H2O2]: 2.94 mm
pH value: 4.5
Temp.: 30°C
[Antibiotic]: 20 mg/l
[H2O2]: 35 mm
Irradiation: 600 mW/cm2
[H2O2]: 15.21 mm
[Dye]: 38 mg/l
[Dye]: 50 mg/l
Temp.: 20°C
[H2O2]: 45.38 mm
pH: 5.7
TOC: 1178 mg/l
COD: 3052 mg/l
[H2O2]: 5.88 mm
UV input power: 560 W
[dye]: 20.0 mg/l
[H2O2]: 116.32 mm
pH: 8.9
UV dosage: 120.70 W/l
[H2O2]: 46.53 mm
Molar ratio of H2O2/Dye: 1588.05
pH: 3.5–5.5
[H2O2]: 50 mm
Molar ratio H2O2/Dye: 1000
[Dye] = 10 μm
[H2O2]: 1.67 mm
pH: 7
Molar ratio H2O2/MO: 587
[Methyl orange]: 7.80 × 10-5 mol/l
[H2O2]: 45.8 mm
Temp.: 28°C
pH value: 3
[H2O2]: 29.41 mm
COD: 5720 mg/l
UV power input: 28.0 W/l
[H2O2]: 116.35 mm
Experimental conditions
COD: 92.3
CR: 98.1
CR: 70
COD: 69
CR: 100
Shu et al. (2006a)
Schrank et al. (2007)
Haji et al. (2011)
AlHamedi et al. (2009)
Aleboyeh et al. (2005)
k: 0.5 min-1
CR: 73
Shu et al. (2005)
Shu et al. (2006)
96.6
CR: 90
TOC: 75
Malik (2004)
Ghodbane and Hamdaoui
(2010)
CR: 84
CR: 100
TOC: 18.9
Ince and Gönenç (1997)
Zuorro et al. (2014)
Kralik et al. (2010)
Jung et al. (2012)
References
COD: 70
TOC: 50
Clofibric acid removal: 99
TOC: 50
Treatment efficiency (%)
276
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catalyst loading, initial pH of solutions, and light intensity
are the important parameters that affect the efficiency of
the TiO2/UV system.
Majority of the previous studies have reported that
degradation efficiency apparently decreases with increasing initial pollutant concentration (Wu 2008). Higher
concentration of organic contaminants inhibits the penetration of light causing fewer photons (H+) that reach
the surface of the catalyst (Liu et al. 2006). This results in
a slower production of hydroxyl radical and reduces the
photo-degradation efficiency. Besides, more organic substances are adsorbed on the surface of TiO2, which also
contributes to the lower formation of hydroxyl radicals.
Wu (2008) investigated the effects of operational parameters on decolorization of C.I. Reactive Red 198 in UV/TiO2based systems. The effects of initial dye concentration on
the rate of dye decolorization were evaluated by changing
the initial dye concentration from 10 to 80 ppm at pH 7.
The decolorization rate constant declined from 0.2424
to 0.0226 min-1 as dye concentration increased. Liu et al.
(2006) also studied the effects of initial dye concentration
of C. I. Acid Yellow 17 on mineralization efficiency. The
result showed that the removal of dye decreased from 70
to 0.2% as the initial concentration increased from 20 to
150 mg/l. In addition, Tang and Huren (1995) and Chun
and Yizhong (1999) found that color removal percentage decreased as the initial concentration increased. The
authors attributed it to shielding of the UV light by the
dye, which causes the light-triggered catalysts to decrease
and reduce the pollutant removal efficiency eventually.
A number of studies have reported that pH value
affects UV/TiO2 system. Dye degradation depends on the
attack by hydroxyl radical, direct oxidation by H+ hole,
and direct reduction by e- in the conducting band depending on the properties of the organic pollutant and the pH
value (Toor et al. 2006). In an acidic medium, H+ holes
react with water molecule producing hydroxyl radical.
Hydroxyl radicals are scavenged at higher pH decreasing
the oxidation of pollutant by hydroxyl radical (Muruganandham and Swaminathan 2006b). The point of zero
charge of TiO2 is at pH 6.8; therefore, below this pH the
surface is positively charged and above that it is the opposite. To investigate the effect of pH on TiO2 performance,
Cho and Zoh (2007) studied the photocatalytic oxidation
of Reactive Red 120. The results showed that acidic solution favored adsorption of negatively charged reactive dye
onto the photocatalyst surface and increased the degradation rate. At lower pH, more hydroxyl radicals will be
generated to oxidize the dye but at higher pH; TiO2 surface
was negatively charged, and the absorption of molecules
on the surface of catalyst decreased. Adsorption of TiO2
277
particles onto dye molecules may be less favorable due to
the repulsion between negatively charged TiO2 particles
and dye molecule. Similar results were reported by Liu
et al. (2006). The highest removal of C. I. Acid Yellow 17
(70.6%) was achieved at pH 3, while the lowest (44.3%)
was observed at pH 11. The properties of the C.I Acid
Yellow 17, which is an anionic dye, make it easier to be
absorbed on the surface of the catalyst at lower pH value.
Similar results were also reported in other UV/TiO2 reaction systems (Tanaka et al. 2000, Liu et al. 2006).
Most of the studies have reported that the efficiency
of UV/TiO2 increases with TiO2 dosage until an optimum
value after which the degradation efficiency reduces
or becomes constant. This is due to excessive TiO2 that
causes a shadow effect that interferes with transmission
of UV light and hinders formation of electron-hole pairs
(Tang and Huren 1995, Thiruvenkatachari et al. 2007).
Besides, decrease in the number of surface active sites at
higher TiO2 concentration also reduces the degradation
efficiency (Toor et al. 2006). Muruganandham and Swaminathan (2006b) evaluated photocatalytic decolorization
and degradation of reactive orange 4 by UV/TiO2 process.
The effects of TiO2 dosage ranging from 1 to 6 g/l on degradation were investigated. Degradation rate increased from
0.075 to 0.290 mol min-1 as TiO2 dosage increased from 1
to 4 g/l. Further increase of TiO2 dosage did not cause any
changes on the degradation rate. The activity of the catalyst became constant above 4 g/l, which may be caused
by scattering of light, particle aggregation, and screening effects. Besides, Chang et al. (2004) investigated the
effects of TiO2 dosage on degradation of synthetic lignin
wastewater. The result showed that the removal efficiency
increased with TiO2 dosage until an optimum level of
10 g/l, beyond which the efficiency reduced with further
increase of TiO2 dosages due to the shadow effects caused
by the excessive amount of TiO2. At optimum dosage of
10 g/l, 50% of decoloration was achieved within 10 min,
and by extending the time to 960 min, almost 80% color
and dissolved organic compounds (DOC) removal was
observed by the researchers. In addition, the literature
study conducted by Ramesh et al. (2008) showed that,
at a specific light intensity, increase in TiO2 concentration beyond a certain value was found to decrease the
efficiency of the treatment. The authors concluded that
higher concentration of TiO2 would affect the passage of
light through the solution and hence will affect the degree
of absorption of light by the catalyst surface. This result
was in agreement with the other literature results (Rizzo
et al. 2014, Fenoll et al. 2015).
Table 4 illustrates the summary of wastewater treatment using UV/TiO2. Various types of organic matter have
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UV-lamps
9 W lamp (350–400 nm)
UV lamp (6 W, 365 nm)
UV germicidal lamp (253.7 nm, 18 W)
Low-pressure mercury arc bulb (15 W)
UV lamp (400 W, 200–550 nm)
UV lamp (254 nm and 8 W power)
UV tube TLAD Philips 415 W/05 (365 nm)
9 W-lamp (Radium Ralutec, 9 W/78,
350–400 nm).
High-pressure mercury lamp (125 W;
λ > 253 nm)
Medium-pressure (254 nm, 500 W)
Medium-pressure mercury lamp (125 W)
UV-lamp (40 W)
Pollutant
Amoxicillin
Pesticides
Oxidation of organic matter in
drinking water
Cryptosporidium parvum
Petroleum refinery wastewater
Methyl tert-butyl ether (MTBE)
Olive mill wastewater (OMW)
Secondary treated municipal
wastewater
DEP
C.I. Reactive Red 198
Reactive Red 198
Direct Red 23
Table 4: Summary of studies on the removal of pollutant by (UV/TiO2).
100
45
30
180
60
24 H
60
240
45
60
300
90
Temperature: 25°C
TiO2 concentration: 0.5 g/l
pH: 5
TiO2 suspension: 1.5 g/l
pH: 6
Nano-TiO2 film
Light intensity: 2–3 mW/cm2
TiO2 suspension: 0.001 g/l
pH: 6–8
[TiO2]: 0.1 g/l
pH value: 3
Temp.: 318 K
TiO2 suspension: 2–3 g/l
pH: 2
[MTBE]: 1 mm
TiO2 suspension: 1 g/l
Filters: 0.45 μm
TiO2 suspension: 1 g/l
pH: 6
TiO2 suspension: 1 g/l
pH: 7
TiO2 suspension: 0.5 g/l
[Dye]: 10–80 mg/l
pH: 4
Millipore filter membrane: 0.45 μm
Temp.: 25°C
TiO2 suspension: 0.3 g/l
[Dye]: 100 mg/l
Temp.: 21–2°C.
TiO2 suspension: 4 g/l
pH: 2
UV-irradiation Experimental conditions
time (min)
Decomposition: > 99
Degradation: 99
TOC: 43–46
Sohrabi and Ghavami (2008)
Kaur and Singh (2007)
Wu (2008)
Mansouri and Bousselmi (2012)
Lydakis-Simantiris et al. (2010)
El Hajjouji et al. (2008)
Hu et al. (2008)
k: 0.0233 min-1
CR: 57
COD: 22
Coliforms removal: 65
Enterococci: 50
78.6%
Saien and Nejati (2007)
Ryu et al. (2008)
Jin-hui (2012)
Chaudhuri et al. (2013)
Dimitrakopoulou et al. (2012)
References
Organic pollutant removal:
90
Log inactivation: 3.3 log10
COD: 25.95
TOC: 8.45
82.14
TOC: 93
Treatment efficiency (%)
278
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UV-lamps
UV lamp (228–420 nm)
40 W-blacklight blue fluorescent lamp
(3.6 W, 368 nm)
UV lamp (6 W, 352 nm)
Medium-pressure mercury lamp (125 W)
Blacklight blue fluorescent lamps (40 W,
254 nm)
UV-black fluorescent lamps (40 W)
UV light (Spectronics, BLE-8T365) was
Pollutant
Procion yellow H-EXL
Reactive Red 120
C.I. Reactive Orange 16
Reactive Red dye 198 (RR dye
198)
Reactive Red 120
Direct Yellow 12
C. I. Acid Yellow 17
(Table 4: Continued)
400
150
90
45
120
90
30
[Dye]: 10 mg/l
pH: 5
TiO2 suspension: 1 g/l
TiO2 suspension: 1.5 g/l
pH: 5
[Reactive Red 120]: 100 mg/l
TiO2 suspension: 2 g/l
pH: 3
TiO2 suspension: 0.3 g/l
Temp.: 25°C
[Dye]: 100 mg/l
pH: 4.6
TiO2 suspension: 2 g/l
[Dye]: 50 mg/l
pH: 5
TiO2 suspension: 2.0 g/l
Millipore filter membrane: 0.45 μm
pH: 4.5
pH value: 3
[Dye]: 50 mg/l
Temp.: 25°C
Flow rate: 0.82 cm/s
UV-irradiation Experimental conditions
time (min)
COD: 73
Liu et al. (2006)
Toor et al. (2006)
Cho and Zoh (2007)
k: 0.0448 min-1
COD: 94
Kaur and Singh (2007)
Kartal and Turhan (2012)
Park et al. (2003)
Barakat (2011)
References
COD: 95
CR: 61
> 99
Degradation: 100
Treatment efficiency (%)
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280
A. Buthiyappan et al.: Degradation and cost of UV IAOP for WT
been successfully treated using TiO2. Based on Table 4, it
is found that optimized value of the initial pH of the solution varies with the targeted organic matter. Besides, TiO2/
UV is mostly used for pollutants of low concentration. It
might be because UV/TiO2 process is energy intensive and
not suitable for high concentrations of pollutants. Retention time of minimum 30 min is required for decolorization or degradation to occur, and it can go up to 400 min
based on the targeted pollutants. Therefore, it is necessary
to optimize the operating parameter for each process to
get maximum removal efficiency. UV/TiO2 systems have
also been successfully applied to achieve COD reduction
of more than 90% and TOC reduction of about 40%.
As a conclusion, the literature shows that heterogeneous photocatalysis is a promising technology for
removal of organic pollutants. It is an environmentally
friendly, low cost, and sustainable process. Apart from
that, the photocatalytic activity and stability of TiO2 are
extraordinary. Nevertheless, full-scale TiO2/UV treatment
systems have not been successfully developed due to low
quantum efficiency, complicated photo-reactor design,
inability to reuse titanium dioxide, generation of intermediate products and by-products, and catalyst deactivation.
It is necessary to resolve all the above stated limitations
to successfully commercialize the system for wastewater
treatment.
2.4 UV/Ozone
Great attention has been given to water and wastewater treatment utilizing ozone as an oxidant (Sharrer and
Summerfelt 2007). Ozone plays a dual role by disinfecting and degrading the target compounds (Summerfelt
2003). Ozone is reported to have successfully degraded
many organic contaminants including pesticides, pharmaceuticals, and municipal wastewater (Beltrán et al.
2012, Bin and Sobera-Madej 2012, Cheng et al. 2013, Lester
et al. 2013b). The mechanism, kinetics of ozone decompositions, and the chain reaction have been extensively
studied (Amat et al. 2005, Sharrer and Summerfelt 2007).
In addition, ozone also can be applied directly in its
gaseous state and therefore does not increase the volume
of wastewater and sludge, and ozone reaction time with
pollutant is short.
The oxidation of organic contaminants takes place
either through direct oxidation by molecular ozone or
indirectly by decomposition of ozone (Bustos et al. 2010).
Direct oxidation of ozone is favored under acidic condition in the presence of radical scavenger, and alkaline
conditions predominate the generation of strong hydroxyl
radical (indirect oxidation) (Liu et al. 2004). At alkaline
conditions, ozone decomposes to secondary, more reactive, and hence less selective oxidants such as OH·, HO2·,
HO3·, and HO4·, are formed (Table 5). Hydroxyl radical is
the main oxidant responsible for indirect oxidation in
ozone decompositions among others (Tezcanli-Guyer
and Ince 2004). Thus, the stability of an ozone solution is
highly dependent on the pH. The depletion rate of ozone
is reduced in alkaline solutions. This is may be due to the
formation of ozonide, O3-, which reacts with H2O2 or OHradicals and reforms ozone.
The important factors that affect organic removal in
ozone systems are dosage of ozone, composition of water,
and reaction-rate constant of ozone with the target contaminant (Huber et al. 2003). Dissolved organic matter
and nitrite are the two most important constituents that
effect the treatment efficiency. Ozone could absorb the
organic matter and improve the particle aggregation by
reducing the electrostatic stabilizing effect of organic
matters. Ozone is decomposed to OH radical through a
radical reaction (refer to Table 5). And it can be promoted
by solutes that could transfer hydroxyl radical into superoxide radical ion. At the same time, the action can be
also inhibited by compounds such as carbonates that do
not promote generation of superoxide radical ion. This
could decrease the rate of reaction due to a drop in the
free radical to complete the degradation process. Besides
that, a radical chain reaction can also be accelerated by
the presence of aromatic compounds in the targeted pollutant which may produce additional hydroxyl radical
reaction between ozone and aromatic compounds. The
Table 5: Mechanism of ozone decompositions at alkaline
­conditions (Tomiyasu et al. 1985).
Reaction
Rate constants
Initial reactions
O3+OH- → HO2-+O2
O3+HO2-→ HO2 +O3-·
Propagation reactions
HO2·→O2-·+H+
O2-·+H+→HO2·
O3+O2-·→O3-·+O2
O3-·+H2O→HO·+O2+OHO3-·+HO·→HO2·+O2-·
O3+HO·→HO2·+O
HO2-+H+→H2O2
H2O2→HO2-+H+
Termination reactions
O3+HO·→O3+OHHO·+H2O2→HO2·+H2O
HO·+HO2-→HO2·+OH-
40.0 m-1 s-1
2.2 × 106 m-1 s-1
7.5 × 105 m-1 s-1
5.0 × 1010 m-1 s-1
1.6 × 109 m-1 s-1
20–30 m-1 s-1
6.0 × 109 m-1 s-1
3.0 × 109 m-1 s-1
5.0 × 109 m-1 s-1
0.25 m-1 s-1
2.5 × 109 m-1 s-1
2.7 × 109 m-1 s-1
7.5 × 109 m-1 s-1
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investigation conducted by Huber et al. (2003) showed
that the rate of reaction is depend on the characteristics of the pollutants, which react slowly with ozone and
rapidly with hydroxyl radical. The reactivity was found to
increase with increases in concentration of organic pollutants and decrease with increases in alkalinity. Lester
et al. (2013a) observed that cyclophosphamide (CPD)
removal occurred mainly through its reaction with ·OH
radicals and decreased with increasing alkalinity and
concentration of peptone (more pronounced) and alginic
acid (less pronounced). The difference in the rate constant of both model compounds influences the removal
efficiency of CPD.
Ozone is can easily absorb UV radiation and generate hydrogen peroxide (intermediate), which then decomposes to hydroxyl radical (Tezcanli-Guyer and Ince 2004).
And it was found that photolysis of ozone yields more radicals than the UV/H2O2 process (Shu and Chang 2005b). The
integration of ozone and UV has been proven to effectively
treat many types of pollutants. Molecular ozone reacts
slowly and selectively with the organic or inorganic compounds compared to hydroxyl radical. There are a number
of obvious differences that could be observed between the
ozone and hydroxyl radical with respect to the reaction
with organic matters. The rate of ozone consumption is
found to decrease rapidly with increasing ozonation. This
is because of mass transfer limitations and self-decomposition of ozone at higher concentrations. However, in
alkaline conditions, high concentration causes rapid production of highly oxidative free radicals, i.e., HO rapidly
reacts with the contaminant, and thus it increases the
rate of reaction because of the continuous availability of
hydroxyl radical within the system, as clear from Table 5.
On the other hand, ozone is selective in nature and attacks
conjugated double bonds of the organic compounds,
which results in formation of other by-products. Thus, it
increases the biodegradability of the wastewater effluent. However, hydroxyl radical follows hydrogen abstraction, electrophilic addition, electron transfer, and radical
chain mechanism. In this way, high reaction rates can be
achieved, which is one of the advantages of using ozone
under alkaline condition.
Ozonation appears to be a more efficient technique
when it is implemented with UV to treat highly polluted
wastewater since it produces additional hydroxyl radical
and hydrogen peroxide via photolysis. A number of studies
have reported that ozone-assisted UV removes organic
contaminants more effectively compared to treatment
with ozone alone (Trapido et al. 1997, Vogna et al. 2004,
Rosenfeldt et al. 2006). UV photolysis of ozone generates
H2O2, which produces hydroxyl radical that reacts with UV
281
radiation as shown in reaction Eqs. (24)–(26) (Gracia et al.
1996, Gong et al. 2008).
O3 + hv + H 2O → H 2O2 + O2 (24)
H 2O2 + hv-→ 2 HO⋅ (25)
2O3 + H 2O2 → 2 HO⋅+3O2 (26)
Ozone/UV has been applied widely in wastewater
treatment even though it is known as a most complex
system. It can degrade organic matter in different ways:
direct ozonation, photolysis reaction, and hydroxyl radiation. Table 6 summarizes the application of O3/UV for
treating different types of wastewater together with their
experimental conditions and treatment efficiency. Based
on Table 6, it is clear that initial pH of the solution, initial
concentration of pollutants, flow rate of ozonation, and
UV radiation play a crucial role in enhancing the treatment efficiency.
Gong et al. (2008) compared the efficiency of O3
and UV/O3 techniques by treating bio-treated municipal
wastewater. The result showed that ozonation alone and
ozonation with UV radiation were both effective in removing UV-absorbing organic substances from wastewater.
However, UV/O3 was found to be much more effective for
mineralization by achieving DOC reduction of 90% compared to ozone alone, which achieved only 36% reduction. In addition, Shang et al. (2007) evaluated oxidation
of methyl methacrylate (MMA) from semiconductor wastewater by O3 and O3/UV processes. The authors reported
that COD and methyl methacrylate removal were 51% and
96%, respectively, in 120 min. The removal efficiency of
MMA by the ozone/UV process was higher than that by
ozonation alone, which only achieved 24% COD removal.
It was because UV radiation could enhance mineralization by decomposing ozone and generating additional
hydroxyl radical simultaneously. Irmak et al. (2005)
studied the decomposition and degradation of two endocrine disrupters, E2 and bisphenol A (BPA) in aqueous
medium by using ozone and O3/UV. In the study, different
O3 dosages, lower and upper level, were used for complete
oxidation of E2 and BPA. The efficiency was determined
based on the initial conversion and complete degradation
of the substrate at initial concentration of 0.40 mm. The
result indicated that the reaction between BPA and O3 was
slower than the reaction between E2 and O3. It was noted
that UV coupled with O3 decreased the O3 consumption
by 22.5% in converting the same amount of E2, within the
limits of the O3 dosages used. The intermediate products
formed in the reaction were analyzed by LC-MS and determined to be the oxidation product of E2 via addition of
O3/·OH radical to different positions of aromatic ring of E2.
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C-18 ultra aqua column employing UV
detection at a wavelength of 210 nm.
Low-pressure mercury UV lamp (wavelength
254 nm)
UV-C lamp (253.7 nm, 15 W)
Low-pressure Hg arc-UV lamp (3 W)
Mercury lamp (125 W, UV-C at 254 nm)
Special low-pressure mercury-vapor lamp
(254 nm)
Low-pressure mercury lamps ( 254 nm, 2.2 W)
Low-pressure mercury lamp (15 W, 254 nm)
Low-pressure mercury lamps
Mercury lamp (125 W, UV-C at 254 nm)
UV Logic model 02AM15
High-pressure mercury vapor lamp (185 nm)
2-propanol
Aniline
Winery wastewater
DEP
Phenol
Ethylbenzene and
chlorobenzene
Tert-butyl formate and its
intermediates
Board paper industry
Sulfamethoxazole
Phenol
Bacteria inactivation in a
freshwater
Olive oil mills wastewater
Low-pressure portable UV lamp (254 nm,
8–32 W)
9H
Low-pressure mercury UV lamp (254 nm)
Biotreated municipal
wastewater
Rhodamine B
60
Low-pressure mercury lamps (254 nm)
MMA
15
60
180
60
150
70 s
60
50
300
90
200
150
120
90
Mercury lamp (125 W, UV-C at 254 nm)
Organic aqueous solution
UV-irradiation
time (min)
UV-lamps
Pollutant
Table 6: Summary of studies on the removal of pollutant by (UV/O3).
TOC: 44.3
k (ozone): 7.86 × 10-3 g/min
pH value: 3–12
[Ozone]: 10 mg/l
[IUV]: 35.96 W/m2
Temp.: 25°C
[Ozone]: 9.10 mg/l
Air flow rate: 150 l/h
pH value: 7
[IPA] = 1000 mg/l,
Ozone dosage = 18.4 mg/min
Temp. = 25°C
[Aniline]: 0.04 mm
[Ozone]: 0.5–0.85 mg/l
Natural pH: 10
Ozone: 0.68 g/min
Ozone dosage: 1.5 and 4 mg/l min
Flow rate: 0.5 l/min
pH: 11
k (ozone): 7.86 × 10-3 g/min
[Pollutant]: 100 mg/l
Ozone dosage: 50–450 mg/l
Influent [ozone]: 39 mg/l
Effluent [ozone]: 15 mg/l
pH: 7
Flows of ozone: 0.8–8 g/h
pH value: 9
[Ozone] = 1.0 mg/l
pH: 7
[SMX] = 1.0 mg/l
pH: 11
k (ozone): 7.86 × 10–3 g/min
Ozone dosages: 0.1–0.2 min mg/l
UV dosage: 50 mJ/cm2
Temp.: 10, 20, 30 and 40°C
pH: 5, 7 and 7
Ozone partial pressure: 0.54, 0.9
and 1.67 kPa
pH: 3
Ozone dosage: 10–40 ml/min
Kusic et al. (2006a)
TOC: > 40
Phenol: Complete
Degradation: < 70
Liu et al. (2012)
k: (0.60–3.38)±0.13 ×
105 m-1 s-1
CR: 97.76
COD: 39.72
82.4–97.5
Cuiping et al. (2011)
Sharrer and Summerfelt
(2007)
Benitez et al. (1997)
Kusic et al. (2006a)
Amat et al. (2005)
k: 184 min-1
TOC: > 40
Phenol: Complete
Bacteria removal: 100
Garoma et al. (2008)
TOC: 99
Cheng et al. (2013)
Oh et al. (2006)
Lucas et al. (2010)
Zhao et al. (2013)
Wu et al. (2008b)
Gong et al. (2008)
Shang et al. (2007)
Kusic et al. (2006a)
References
Degradation: 50
TOC: 26
Degradation: 100
TOC: 23
COD: 85
COD: 51
Treatment efficiency (%)
Experimental conditions
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Based on the results obtained, it is clear that the degradation efficiency depends on the types of pollutants and the
dosage of ozone used.
Lucas et al. (2010) used O3/UV at the natural pH
to treat winery wastewater of initial pH = 4, 7, and 10, to
investigate the effect of radical formation on the rate of
COD removal. The authors observed the fastest COD
removal at pH 10 as a result of fast reaction between the
organic matter and the radical species (i.e., ·OH, HO2·, O2·-,
and O3·-). Ozone self-decomposition to radicals occurs
at alkaline pH due to initiation reaction (Tomiyasu et al.
1985). Besides, a sharp decrease in the pH of the solution
showed that strong oxidation took place. The reduction in
pH was a result of the formation of dicarboxylic acids, CO2,
and carbonic acids. The result obtained by Arslan et al.
(2014) was in agreement with that reported by the previous researchers. In this study, the experimental ranges of
the factors were selected as initial pH of 2–11, ozone concentration of 0–15 mg/l, and reaction time of 20–60 min.
The authors observed that O3/UV yielded the highest degradation efficiency of raw hospital wastewater at initial
pH of 8.0 and O3 concentration of 4.2 mg/l within 27 min.
Kusic et al. (2006a,b) discussed the effect of UV light, pH,
and ozone dosage on the degradation of phenol in UV/O3
systems. The highest calculated rate constant for phenol
degradation, k = 0.1936 min-1, was obtained. The highest
phenol degradation rates were observed in strong alkaline media. The result showed that degradation efficiency
increased with pH up to pH 8 and then it reduced at pH
9 and 10. However, the maximum TOC removal of 44.3%
was observed at pH 11. In addition, Gurol and Vatistas
(1987) also demonstrated that based on their experimental results, molecular ozone was dominant only at low
pH level (acidic). Free radicals such as OH radicals were
the main influencing factor at neutral or basic pH. Such
observation supports that UV could help enhance the degradation efficiency by generating additional free radicals.
Besides, ozone concentration can also significantly
influence the oxidation process. Ozone molecules should
dissolve in the aqueous solution and diffuse into the
system for continuous oxidation process for an efficient
oxidation process (Gurol and Vatistas 1987). It was proven
by Amat et al. (2005) in their control experiment. The
result showed that the COD and TOC removal efficiency
under UV light treatment alone was much lower than
that under the combination of ozone and UV. The ozone/
UV combination has shown significant synergetic effect.
Beltrán et al. (1997) observed that O3/UV radiation was the
best oxidation method to remove COD and TOCD regardless of wastewater type. UV light intensity is another
operating parameter that can influence oxidation rates.
283
Wu et al. (2008a,b) observed that ozonation process was
strongly enhanced by the presence of UV light for degradation of 2-propanol. The enhanced degradation is due to
abundant generation of hydroxyl radicals. Ozone absorbs
UV radiation and generates hydrogen peroxide. The subsequent photolysis of hydrogen peroxide generates hydroxyl
radicals as shown in Eq. (24). Hydrogen peroxide could
quicken ozone decomposition into OH radicals. Their
study showed that OH radicals play an important role as
an active species in the photolytic ozonation process.
Considering the aforementioned discussion, the combination of UV with O3, at optimum design and optimized
conditions, will yield higher degradation and removal
rate compared to using ozonation alone regardless of
type of wastewater. The synergistic effects of this combination have been discussed in the literature. It is believed
that the synergistic effects are attributed to intensive generation of highly oxidative and non-selective reagents
such as hydroxyl radicals. Based on Table 6, to date,
COD removal of more than 90% has been recorded for
the treatment system utilizing UV with O3. Besides, the
simplicity of the process makes it a suitable alternative
compared to the other oxidation methods. Choosing a
proper irradiation source, initial pH value, and optimum
concentration and developing new operative designs are
important to make the process more successful. Among
them, an appropriate reactor design could considerably
stimulate the synergistic effects of the system for industrial scale-up.
2.5 UV/Persulfate
More recently, sulfate radical (SO4·-)-based AOPs have
been identified as a promising technique for degradation of recalcitrant pollutants. Sulfate radical is one
of the strongest aqueous oxidizing species with a high
redox potential of 2.5–3.1 V (Cai et al. 2014) similar to
that of hydroxyl radical which is of 2.8 V. Sulfate radicals
also offer several advantages over other oxidants such
as longer half-life, fast kinetics, higher stability than
hydroxyl radical, greater transport distances in the subsurface level, and ability to work in a wide range of pH,
and it can be activated by low-cost oxidant precursors.
However, although it is more stable than hydroxyl radical,
the narrow selectivity of sulfate radical toward organic
matters makes it less efficient compared to hydroxyl
radical. Sulfate radical reacts through electron transfer or addition or hydrogen abstraction. The high redox
potential of sulfate radical causes it to produce radicals
from many anions through electron transfer (Criquet and
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A. Buthiyappan et al.: Degradation and cost of UV IAOP for WT
Leitner 2009). These attributes collectively make sulfate a
viable option for chemical oxidation of a broad range of
contaminants.
Persulfate anions with a redox potential of 2.1 V are
widely used as sulfate-based precursor oxidants, and
they demonstrate a great ability in degradation of refractory compound (Watts and Teel 2006, Rastogi et al. 2009,
Lin et al. 2014). Peroxydisulfate (PDS, S2O82-) and peroxymonosulfate (PMS, HSO5-) are two types of oxidants that
have been used for degradation of synthetic dyes, pharmaceuticals, chlorophenols, PAHs, and pesticides. PDS
is commonly used as a source of sulfate radical because
of its high aqueous solubility and stability at room temperature and relatively low cost (Cai et al. 2014). Besides,
persulfate reduction is relatively harmless and considered environmentally friendly (Cai et al. 2014). The commonly used salts as a source of sulfate radicals for organic
matter decompositions are sodium persulfate (Na2S2O8),
potassium persulfate (K2S2O8), and ammonium persulfate
(NH4)2S2O8 (Huang et al. 2005). It has been reported in
the previous work that potassium persulfate yields better
organic matter removal efficiency at its natural pH, and it
is much cheaper compared to other salts.
Although it is a strong oxidant, the reactivity of persulfate ions is mostly slow at room temperature. It has been
postulated that persulfate anion (S2O82-) can be activated
by heat, UV, electron transfer, and transition metals to
generate a stronger oxidant sulfate radical (SO4·-). (Liang
et al. 2003, Liang et al. 2004a,b, Huang et al. 2005, Criquet
and Leitner 2009, Mora et al. 2009) [Eqs. (27)–(38)]. The
formed radical has a redox potential of 2.6 V and is kinetically improved.
In accordance with Eqs. (1)–(15), in the reaction accelerated by UV light, the oxidation process is begun by
generation of the sulfate radicals followed by hydroxyl
radical. Photolysis of S2O82- results in the formation of two
SO4·- radicals [Eq. (1)] (Tsao and Wilmarth 1959). It is noted
that SO4·- has a maximum optical absorption spectrum at
440–450 nm and a molar extinction coefficient between
460 and 1600 m-1 cm-1. The formed radicals are powerful
oxidizing agents which attack the recalcitrant compounds
in the contaminated water and cause a complete decomposition of those compounds into carbon dioxide and
water. Although sulfate ion is formed as an end product, it
is inert and not considered as a secondary pollutant. The
sulfate ions formed help decrease the pH and increase the
salt content of the effluents.
S2O8 2- + hv→ 2SO4⋅- (27)
SO4⋅- + RH 2 → SO4 2- + H + + RH ⋅ (28)
RH ⋅ + S2O8 2- → R + SO4 2- + H + + SO4⋅- (29)
SO4⋅- + RH → R⋅ + SO4 2- + H + (30)
2R⋅ → RR ( dimer ) (31)
SO4⋅- + H 2O → HSO4- + OH ⋅ (32)
HSO4- → H + + SO4 2- (33)
OH ⋅ + S2O8 2- → HSO4- + SO4⋅- + ½O2 (34)
SO4⋅- + OH ⋅ → HSO4- + ½O2 (35)
2OH ⋅ → H 2O2 (36)
OH ⋅+ H 2O2 → H 2O + HO2⋅ (37)
S2O8 2- + H 2O2 → 2H + + SO4 2- + O2 (38)
S2O8 2- + e → SO4⋅- + SO4 2- (39)
S2O8 2- + heat → 2SO4⋅- (40)
S2O8 2- + M n+ → SO4⋅- + SO4 2- + M ( n+1 )+ (41)
(Mn+ (Transition metals) = Ag+, Cu+, Co2+, Fe2+, Mn2+)
R = organic compounds
Among the AOPs, the homogeneous AOPs employing
PDS and UV/PDS have been found to be very effective in
degrading refractory pollutants. Regardless of the type
of precursor used, persulfate oxidation is highly sensitive to the process conditions. Therefore, the optimization of operating parameters is very important in order to
determine the influence of each parameter on the process
efficiency. Based on the literature study, it is noted that
operating parameters including initial pH, initial concentration of pollutant, initial concentration of persulfate,
and temperature have an impact on the system efficiency.
Khataee and Mirzajani (2010) are the pioneer authors
who have investigated the photo-oxidative decolorization
of the textile dye C.I. Basic Blue 3 (BB3) through UV/PDS
process. In their work, they reported the effects of operating parameters on photochemical treatment of a dye solution. The decolorization efficiency was investigated in the
presence of potassium PDS (K2S2O8), irradiated by a 30 W
UV-C lamp. The result showed that the decolorization efficiency was affected by operating parameters such as the
reaction time, UV light intensity, initial concentration of
BB3, and amount of PDS.
Among other parameters, initial concentration of
PDS is very important as it supplies free active radicals
for decolorization and degradation process. The proper
addition of PDS is necessary to improve the degradation
rate. At higher concentration of PDS ion, more sulfate
and hydroxyl radicals are available to attack the aromatic
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A. Buthiyappan et al.: Degradation and cost of UV IAOP for WT
rings, and thus the decolorization efficiency increases.
Khataee and Mirzajani (2010) reported that the initial
concentration of PDS was found to be an important
parameter for the photooxidative decolorization of BB3
in the UV/S2O82- process. It was noted that the increase of
initial S2O82- concentration from 0.1 to 1.8 mm increased
the pseudo-first-order rate constant from 0.0036 to 0.1933
min-1. However, once it exceeded a certain value (1.8 mm),
a decrease in the photodegradation rate was observed. It
is because as the concentration of S2O82- increases, more
sulfate and hydroxyl radicals are available to attack the
aromatic rings and thus the rate of reaction increases.
However, at higher S2O82- amounts, excessive amount of
hydroxyl radicals are generated, which will recombine to
form less reactive H2O2 [Eq. (10)] to become scavenger of
OH· radical as well as the sulphate radical. This observation was in agreement with the result of the other studied
reactions (Khataee 2010, Yoon et al. 2011).
In addition, initial pH of the solution also plays an
important role in degrading pollutants. Saien et al. (2011)
reported that the efficiency of dye decolorization generally
decreased mildly with either an increase or decrease relative to the system’s natural pH of 6.0. It was because SO4·species may undergo reactions with H2O or -OH to generate
·OH under neutral or alkaline pH conditions, according to
Eqs. (16) and (17). Presence of SO42- anion can inhibit the
reactivity of ·OH or SO4·-. HSO4- is with pKa value of 2.0,
and above this SO42- is the predominant species rather
than HSO4-. Therefore, they cause hydroxyl radical scavenging and slow down the decolorization efficiency.
SO4⋅- + H 2O → H + + SO4 2- +⋅ OH (42)
SO4⋅- + - OH → SO4 2- +⋅ OH (43)
However, in an acidic media, additional SO4·- could be
formed with acid catalyzation, according to Eqs. (18) and
(19). Higher generation of SO4·- could also cause radical
scavenging reaction instead of reaction with organic
matter. However, the researchers did not observe any variation in the degradation efficiency in the cases of acidic
initial pH values, and they attributed it to less generation
of H+ during the reaction. Therefore, in this study, the
natural initial pH of 6.0 is considered an optimum pH with
the highest discoloration efficiency obtained.
S2O8 - + H + → HS2O8 - (44)
HS2O8 - → SO4⋅- + SO4 2- + H + (45)
Similar trends were also reported by Criquet and
Leitner (2009). The authors investigated the degradation
of acetic acid by UV/PDS. They reported that the reaction rate constant decreased with increasing pH. In an
285
alkaline solution, OH- acts as a scavenger of sulfate radicals to slow down the degradation reaction by forming
hydroxyl radical through reaction with sulfate radicals.
This decreases the reaction rates. At acidic conditions,
additional sulfate radicals can be formed from acid catalyzation, which aids in the degradation process as shown
in Eqs. (42) and (43).
The influence of pollutant concentration on degradation is very important from the mechanistic and application point of views. The effect of initial concentration
of dye on decolorization of C.I. basic blue 3 was investigated by Khataee and Mirzajani (2010). They observed a
decreased decolorization efficiency with increase in initial
BB3 concentration. This is because increase in the concentration causes a rise in the internal optical density, and
the solution becomes more impermeable to UV radiation.
As a result, lesser amount of UV radiation reaches PDS
to form radicals, and degradation efficiency decreases.
Besides, higher concentration of pollutant leads to formation of more radical scavengers, which causes competition
among the free radical scavengers and carbonaceous substances. It is important to determine the chemicals such as
hydrogen peroxide based on the COD or concentration of
pollutant. Their finding was in agreement with the finding
of Salari et al. (2009) that photooxidation efficiency
decreased as initial dye concentration of C.I. Basic Yellow
decreased. The initial dye concentration increased at the
same concentration of S2O82- by using UV/S2O82- process in
a rectangular continuous photoreactor. Besides, the study
conducted by Yoon et al. (2011) on methylated arsenic
species using UV/S2O82- process showed that pH had an
impact on degradation efficiency. They observed that the
oxidation efficiency of arsenic species was the highest at
pHi = 3 than pHi = 7 and pHi = 10.
In addition, removal efficiency of the pollutant is
also affected by UV light intensity. The study conducted
by Salari et al. (2009) showed that no noticeable color
removal was observed when the irradiation was applied
in the absence of S2O82-. However, significant dye degradation was noticed by the authors using S2O82- with the presence of UV radiation. This is due to formation of hydroxyl
radical during the reaction. Khataee and Mirzajani (2010)
also observed that removal rate increased from 0.0769 to
0.2243 min-1 with increase in UV irradiation intensity from
9.5 to 33.1 W/m2. It is believed that increase in UV radiation
enhances the production of sulfate and hydroxyl radicals,
which further increases the removal efficiency.
Based on the above discussion, it should be noted
that UV/S2O82- process is very sensitive to the operating parameters such as initial concentration of pollutants, initial concentration S2O82-, pH, and UV radiation.
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Therefore, it is necessary to optimize the parameters to
enhance the removal efficiency of each reaction. Besides,
due to the complexity of the reaction involved in the UV/
S2O82- process, researchers should focus more on the kinetics involved in the reactions. The kinetics study can also
be used to identify the intermediates formed, and it is
helpful for toxicity determination. Since the photochemical system is quite complex and it involves radiant energy
balance, the spatial distribution, mass transfer, and
mechanisms of a photochemical degradation involving
radical species such as hydroxyl radical, a proper mathematic modeling, or simulation technique is important in
order to accurately predict the system efficiency. Furthermore, mechanisms and chemical phenomena are useful
for scaling up the photochemical systems.
2.6 Various combinations of UV-AOPs
It can be noted that combination of UV radiation with different techniques could enhance the treatment efficiency.
But appropriate techniques must be combined together to
give economically and technically feasible options. Therefore, a few factors must be taken into consideration when
combining the system such as the cost of the treatment
system, flexibility of the treatment, type of pollutants, and
biodegradability of the wastewater.
Based on the previous studies conducted, the combined systems have been found to significantly improve
the efficiency of treatment. Table 7 summarizes some of the
work done using those processes with the type of equipment along with experimental conditions and important
results obtained in the work. The efficiency of wastewater
treatment system is commonly measured by the COD, TOC,
biochemical oxygen demand (BOD), dissolved organic
carbon, BOD5/COD ratio, and concentration of specified
pollutants by using HPLC, GC-MS, and ion chromatography to confirm the oxidation of inorganic species (such
as CI- and NO3-), as well as toxicity analysis and measurement of decolorization by UV-spectrophotometric method.
The main driving mechanism of a treatment system is the
generation of a free radical through the chemical reaction
which can increase the rate of reaction. The combination of
UV-radiation with H2O2, O3, and other catalysts is promising to increase the production of free radical, and recently
few studies have been attempted to study the efficiency of
using the combined system (Lester et al. 2011).
In a recent study, Zhang et al. (2013) have reported the
efficiency of using combined UV/TiO2/H2O2 under high UV
photon flux irradiation for decolorization of methylene
blue (MB). The photodegradation of dye using UV/TiO2/
H2O2 process was much more effective than UV/TiO2 process
alone. UV/TiO2/H2O2 process achieved 98% decolorization of 20 mg/l MB under optimal conditions in 10 s. The
authors have observed that the addition of H2O2 enhanced
the photocatalytical reaction rate constant by almost three
times. The photonic efficiency calculated based on the
various experimental conditions showed that H2O2 could
improve the light utilization efficiency of photocatalytic
process. Experiments with different dosages of H2O2 were
carried out to investigate the effect of H2O2 dosage on the
photocatlytical process. The reaction rate constant showed
remarkable increase with the concentration of H2O2 up to
100 mg/l; a very small increase was observed at H2O2 concentration of 200 mg/l, and no changes in rate constant
was observed beyond 200 mg/l. since hydrogen peroxide is
one of the prime factors contributing to the cost of photocatalytical process, it is important to minimize the amount,
so authors have used H2O2 concentration of 100 mg/l as an
optimum in this study.
Kuo et al. (2013) studied the decolorization of C.I.
Reactive Red using TiO2/powered activated carbon (PAC)
under UV and visible light irradiation. The authors have
studied the effects of the C/Ti ratio, calcination temperature, photocatalyst dose, CI reactive red 2 (RR2) concentration, wavelength of light, and pH on the decolorization
efficiency of RR2 by TiO2/PAC/UV. Based on the result
obtained, it is confirmed that PAC increased the photocatalytic efficiency of PAC/TiO2/UV than TiO2/UV by acting
as a good adsorbent and photo-induced electron acceptor. The TOC removal efficiency utilizing UV (254 nm)/TiO2
and UV (254 nm)/TiO2/PAC systems were 40% and 56%,
respectively, after 1 g/l photocatalyst was added, and the
reaction time was fixed at 240 min. Result showed that
efficiencies of TiO2/PAC in both decolorization and TOC
removal were better than using TiO2.
Kim et al. (2011) applied microwave/UV/O3/H2O2/TiO2
photocatalyst hybrid process system to investigate the
photocatalytic decomposition characteristics of three different single-component organic dyes and their mixture.
The authors have evaluated different combinations of
microwave, ozone, hydrogen peroxide, and UV to find the
optimal value to enhance the photocatalytic decomposition efficiency of organic dyes. In this study, microwave
irradiation was added to accelerate the decomposition
reaction by activating pollutants and photocatalysis.
Moreover, oxidants such as ozone and hydrogen peroxide
were added to enhance the decomposition efficiency of
the system. When microwave irradiation was used alone
with photocatalyst, the effect was not significant, and
remarkable increase of decomposition rate was observed
when it was used together with other auxiliary oxidants.
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Amoxicillin,
ampicillin and
cloxacillin antibiotics
Real textile effluents
Disperse blue 1
UV/TiO2/H2O2
UV/TiO2/H2O2
UV/TiO2/H2O2
C.I. Reactive Red 2
MB
High photon flux/
UV/TiO2/H2O2
UV/TiO2/PAC
Pollutant
Method
– Irradiation source was a highpressure lamp (250 W) assembled
with a reflector, shutter, timer and
air cooler in a closed box
– The UV photon flux was adjusted
by transmission filters of different
density
– The source of UV irradiation was a UV
lamp with a nominal power of 6 W,
emitting radiation at 365 nm and it
was placed above the reactor
– 600 ml reactor
– The lamps were vertically fixed onto
the top wall of a wooden box (80
cm × 80 cm × 50 cm)
– Four fans were placed in different
positions on the side walls of the
reactor to minimize the heating
effect produced by the lamps
– The internal walls were covered with
aluminum foil to avoid radiation
– Photochemical reactor made of Pyrex
glass equipped with a magnetic
stirring bar, water circulating jacket
and an opening for supply of air
bubble was used
– Irradiations were carried out using
a 125 W medium-pressure mercury
lamp
– An 8 W lamp (254, 365, or 410 nm)
was placed inside a quartz tube as
the light source
– The reaction temperature was 25°C
in all experiments
Equipment details
– pH was 7
– The optimal calcination temperature
for forming TiO2 – TiO2/PAC = 400°C;
surface areas of TiO2 = 42 m2/g
– TiO2/PAC = 108 m2/g
– Irradiation time: 60 min
– [Dye]: 0.25 mm
– [H2O2]: 0.3 ml
– TiO2: 1 g/l
– V: 250 ml
– pH: 3.0
– TiO2: 0.25 g/l
– H2O2: 10 mm
– Fe2+: 3.5971 × 10-2 mm
– TiO2: 1.0 g/l
– pH∼5
– Initial COD: 520 mg/l
– [H2O2]: 2.9412 mm
–U
V photon flux ranging:
3.13 × 10-8 to 3.13 × 10-6 einstein cm2/s
– H2O2 dosages: 5.8824 mm
Optimum operating conditions
Table 7: Summary of studies on the removal of pollutant by various combination UV-AOPs.
– Percentage of sanatase increased
with the amount of PAC in TiO2/PAC
– The spectra indicate that C doping
of TiO2 shifted the absorption edge
from 418 nm to a longer wavelength
of 471 nm
– TOC removal percentage in the
UV(254 nm)/TiO2 and UV (254 nm)/
TiO2/PAC systems was 40 and 56%,
respectively
– Photocatalytic reactions
approximately followed a pseudofirst-order kinetics
– Complete degradation of amoxicillin,
ampicillin and cloxacillin in 30 min
– Fenton reactions based treatment
proved to be slower and exhibited
more complicated kinetics than the
ones using TiO2
– UV/TiO2/H2O2 reached reduction
levels higher than 90%
– The use of only peroxide or Fenton
reagent resulted in COD reductions
of 60% and 80%
– More efficient generation of hydroxyl
radical and inhibition of electron/
hole pair recombination
– Optimum catalyst concentration has
to be found to avoid excess catalyst
and ensure total absorption of
efficient photons
– H2O2 could reduce energy
consumption remarkably
– 98% decolorization of 20 mg/l MB
could be achieved in 10 s
Important results
Kuo et al. (2013)
Saquib et al.
(2008)
Garcia et al. (2007)
Elmolla and
Chaudhuri (2010)
Zhang et al. (2013)
References
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Pollutant
Humic
BPA
Dyeing wastewater
Clofibric acid
Method
UV/TiO2/oxidants
UV/TiO2/zeolitebased systems
TiO2/H2O2/UV
VUV/H2O2
(Table 7: Continued)
– Cylindrical glass reactor (inner
diameter 8.0 cm, height 25 cm,
volume 800 ml) with tap water
running through a jacket with
temperature controlled at 10°C.
– The VUV lamp (10 W, combining 90%
UV254 and 10% UV185)
– The UV chamber consisted of 16 UV
lamps
– The effective volume: 1.88 l,
– Total system volume (reactor plus
mixing tank and piping): 10 l
– The system was composed of a 64 W
UV lamp, pH meterand an oxygen/
ozone generation device
– The working volume of the
photocatalytic reactor was 3.1 l
– A maximum light intensity output of
254 nm
–The photoreactor consisted of four
30 cm-long quartz tubesand five
lamps
– The external diameter of each tube
was 1.2 cm and the internal diameter
was 1 cm
– The column photoreactor was
irradiated with 254 and 365 nmUV(8
W, Philips) at room temperature
– The intensities of 254 and 365 nm
UV were 4.44 and 0.7 mW/cm2,
respectively
Equipment details
[C]0 = 10.0 mg/l; [H2O2]0 = 2.9412 mm
T = 10°C.
TiO2 dosage = 1.82 g/l
[H2O2] = 28.824 mm
Reaction time = 20 min
– The flow rate of the BPA solution
was1.4 ml/min
–P
hotocatalyst dosage in each quartz
tube was 6.5±0.5 g/l
– The optimal experimental pH for
the degradation of BPA in the TiO2
system was 6
[TiO2] dosage = 0.3 g/l
[H2O2] = 1.4706 mm
O3: 20 g/m
[K2S2O8] = 50 mg/l
Optimum operating conditions
– Removal of humic acid and
hazardous heavy metals was much
greater when H2O2 was used as the
oxidant
– The SO2- and OH radical produced
are responsible for the rapid
photodegradation
– Supporting TiO2 on zeolite increased
its adsorption capacity
– Reduced the extent of
photogeneratedelectron-hole
recombination in TiO2, enhancing
photocatalytic activity
– The deposition of Cu2O on TZ
modified its photoconductive
properties by increasing the charge
separation efficiency between
electrons and holes
– Cu2O also acted as a trap
for electrons, inhibiting the
recombinationof electrons and
holes
– The removal of TOC increased with
increasing concentration of FeCl3
upto600 mg/l.
– Maximum removal efficiency (85%)
of dye wastewater was obtained at
the optimal value of TiO2 and H2O2
CA removal efficiency reached over
99% after 40 min
Important results
Li et al. (2011a)
Lee et al. (2005)
Kuo et al. (2014)
Jung et al. (2009)
References
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So it suggests that microwaves produce active species to
be used for degradation when it combines with auxiliary
oxidants. The authors have also mentioned that the effect
of H2O2 is prominent since further addition of H2O2 above
the optimum hampered the reaction. This is due to the
blocking of surface sites by H2O2 and the OH radical scavenging by H2O2.
Recently, Patel et al. (2013) have studied the photochemical decolorization and mineralization of Reactive
red 241 (RR241) in aqueous solution using a photo-Fenton
process and its combination with activated charcoal and
titanium dioxide. According to the COD and TOC removal
efficiency, authors suggested that the UV/Fenton/TiO2 was
more effective than homogeneous photo-Fenton and UV/
Fenton/activated carbon (AC). UV/Fenton/TiO2 process
required 120 min for complete decolorization and yielded
a maximum TOC reduction of 43.9% after 240 min in contrast with UV/Fenton/AC, which only gave 14.9% of TOC
reduction after 240 min. Photo-Fenton/TiO2 is more effective than other processes possibly due to the fact that
absorption of UV radiation with TiO2 generates conduction band (CB) electrons and valence band (VB) holes.
These electrons (e) may react with Fe3+ to convert Fe2+
rapidly and increase the concentration of Fe2+.
The preceding discussion shows that a number of
studies have been directed to overcome the limitation of
existing UV-based single process AOPS. Recently, AOPs by
adding more than one oxidants and catalyst with UV radiation is gaining great attention from researchers. Based on
previous studies, it was noted that a combination of AOPS
may give advantages such as good photochemical efficiency, increase in the surface area of catalyst, and thus
increases in the incident light absorption for the photocatalyst which produces more OH radical and accelerates auxiliary oxidants to produce more active species for
degradation purposes. However, very limited number of
studies have been conducted to prove the advantages over
other AOPs. So, more precise and quantitative analyses
in this area are required for more reliable evaluation and
application.
3 M
odels for the design and
optimization of UV-AOPs system
Referring to this context, UV-integrated AOPs have been
proven to be efficient for pollutant destruction, and there
are few full-scale applications already developed. Among
others, UV/H2O2 system is the most often studied fullscale system for the destruction of organic compounds
289
by the combined mechanisms of direct UV photolysis and
hydroxyl radical reactions (Kruithof et al. 2007, Swaim
et al. 2008, Audenaert et al. 2011). However, publications on full-scale application of UV-AOPs are still scarce
to date. Even though many researchers have proven the
effectiveness of the technology in small-scale batch of
UV-AOPS for degradation of a variety of pollutants, there
are some factors that stymie the application of UV-AOPs at
large scale. Lack of proper modeling and simulation tools
for predicting and analyzing the system’s performance are
among the major factors hindering their practical implementation (Wols and Hofman-Caris 2012). The modeling
of a system offers advantages such as assessment of the
performance of UV-AOPs based on the optimized value
and to give a great assistance to scale up the systems. The
model being developed to account for varying operating
conditions depends on the properties of the wastewater
and the UV-AOPS system.
Designing a large-scale UV-AOPs system requires
knowledge on system configuration as well as the chemical kinetics (reaction mechanisms and kinetic rate constants). System configurations including reactor design,
pipe and fittings, lamp number, and lamp orientation
are important (Wang et al. 2012). Though many studies
have been conducted on efficiency of UV-AOPs, the existing information is not sufficient for designing a full-scale
treatment system. Therefore, it is important to identify
appropriate numerical tools to design and optimize an
UV-AOPs system which is capable of combining hydrodynamic reactor models, fluence rate models, and chemical kinetic models. As mentioned earlier in the paper,
UV-AOPs systems strongly depend on the composition and
properties of the wastewater, type and concentration of
oxidants, intensity of UV radiation, and type and dosage
of catalyst. Optimization of these parameters is necessary
in order to develop a full-scale system.
CFD is a powerful numerical tool that has recently
been widely used to design UV reactors and assess performance of UV reactors (Verbruggen et al. 2015). CFD is
capable of numerically solving the fluid dynamics equations through space and time, including conservation of
mass, conservation of momentum, and conservation of
energy (Boyjoo et al. 2014). The developed equation can
describe both the physical and chemical changes within
a reactor by combining appropriate boundary and initial
conditions. A CFD model for UV-AOPS also includes the
spatial variations of fluence rate within the UV reactor.
Many researchers have demonstrated the importance of
combining hydrodynamics of UV reactor with fluence
rate models to predict the effectiveness of the degradation
process. The combination allows optimization of lamp
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A. Buthiyappan et al.: Degradation and cost of UV IAOP for WT
placement, minimization of UV screening, and improved
prediction of contaminant removal in different UV reactors. Therefore, an effective CFD simulation must include
turbulence models, fluence rate models, and accurate
reaction mechanisms (direct and indirect photolysis)
describing the oxidation of the contaminant (Verbruggen
et al. 2015). UV-AOPs must incorporate the impact of turbulent mixing into any chemical reactions that occur within
the system during the simulation (Ghafoori et al. 2013).
Standard k-ε turbulence model is often used for investigating the hydraulics within a flow-through reactor. UV
fluence rate depends on the distance of pollutant from the
lamp and is a function of absorptive characteristics of the
media through which the light is irradiated. In AOPs, UV
light is absorbed by organic pollutant, oxidants, and catalyst. Therefore, the transmitted fluence rate is reduced,
and consequently, the generation of hydroxyl radical
throughout the UV reactor is affected.
The reaction pathways involved in the degradation
of pollutants to its intermediates and final products have
been carefully studied in numerous studies. The reaction
pathways are specific to the parent compound and chemical constituents in the water. Such studies have provided
detailed information on oxidation process and numerical kinetic models. Previous works have clearly shown
that the kinetic model process is important for predicting the conversion rate of reactants into the products. For
example, kinetic models based on Langmuir-Hinshelwood
equations can adequately describe reactivity results and
provide kinetic constant and equilibrium adsorption constant for degradation of organic compound. The defined
reaction kinetics of a UV-AOPs system can be combined
with simulation models (CFD) to identify irradiance distribution profiles and hydrodynamic and turbulent characteristics of a UV reactor system. The developed kinetic
reaction, in combination with the simulation model, may
possibly describe the performance of a system based on
the irradiation distribution profiles and hydrodynamic
and turbulent characteristics of an UV reactor system
(Verbruggen et al. 2015).
There is a number of publications focusing on application of CFD models for AOPs to design and optimize
the system based on the degradation of targeted pollutants (Chen et al. 2011, Wang et al. 2012, Boyjoo et al.
2014). However, it is noted that a very limited number of
studies have been performed with CFD to stimulate UV/
AOPs. Alpert et al. (2010) have evaluated the performance
of comprehensive CFD/UV/AOP models for the degradation of an indicator organic contaminant. The CFD results
were validated with pilot-scale experiments. Besides that,
CFD also has been used for modeling a photo-Fenton-like
process by Ghafoori et al. (2013). The developed CFD
model could be used to combine with the kinetic models
to obtain more accurate performance prediction. In this
study, the authors have developed a valid kinetic model
based on the photochemical reactions and rate constants.
The authors observed a good agreement between CFD simulation results and the experimental data which indicate
that the model is accurate. But it should be noted, since
very little information is available on the CFD modeling of
photo-Fenton-like process, that it is difficult to assess the
suitability of this modeling for photo-Fenton application.
Additionally, many researchers have applied CFD for
modeling different photocatalytical reactors (Pareek et al.
2003, Mohseni and Taghipour 2004, Romero-Vargas Castrillón and de Lasa 2007). It showed the efficiency of the
model by saving cost and time when it is coupled with
photocatalytical reactor. Recently, Jatinder and Kumar
have studied the application of CFD in combination with
response surface methodology (RSM) to optimize the
operating parameters and improve the performance of
immobilized titania-based annular photocatalytic reactor
for the removal of Rhodamine B from water. CFD modeling of the photocatalytical process in annular reactor was
designed, and the output of the CFD model was evaluated
with experimental results to validate the model’s predictions. Then, the developed CFD model was used in combination with RSM to optimize the process parameters.
The aforementioned discussion showed that CFD
models can be effective in predicting the degradation
efficiency of UV-AOPs. The combination of turbulence
models, fluence rate models, and kinetic rate equations
is important to stimulate the reaction using CFD. CFD is
also capable of optimizing the energy and chemical usage
to achieve better degradation efficiency. In addition, the
main characteristic of a CFD method is its ability to accurately calculate the spatial variation of flow rate, reaction
rate, and concentrations at the reactive surface. CFD can
also simultaneously estimate several parameters from one
experiment. However, more studies have to be conducted
to evaluate the efficiency of CFD to predict the removal
efficiency of a variety of organic contaminants. To date,
only a few researchers have used CFD to model their
systems from batch to pilot scales.
4 Cost of UV-combined AOPs
AOPs have emerged as a technically feasible treatment
method for various refractory industrial wastewater, but
a proper economical analysis is needed for an industrial
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Capital cost
291
• Design’s flow rate
• Total energy requirement in the AOP reactor
• The energy supplied by single unit of AOP
• The cost of AOPreactor (UV/US/Microwave)
• The cost of the devices used
• The power density used for UVenergy in the
treatability study
Part replacement cost
• Lamp r for UV system
• Transducer for US system
• Ozone generator parts for ozone system
• Catalyst holder for catalytic systems
• Electronic circuit replacements
• Replacement for microwave system
• Replacement of membrane column
Maintenance and
operational cost
Chemical cost
• Hydrogen peroxide
• Iron (II) sulfate
• Sulfuric acid
• Sodium hydroxide
• Catalyst such as TiO2, CuO, CuSO4, ZiO2
Electrical cost
• H2O2 pumping
• Lamp operation
• Heater resistances
• pH-meter
• Compressor
• Syringe pump for H2O2 dosification
• US device
Labor cost and analytical cost
• Water sampling cost
• System inspection, replacement and repair
• Sampling frequency the labor required to conduct
• Cost of chemicals required for analysis
Figure 2: The elements to be included in treatment system cost estimation based on capital, maintenance, and operational cost for UVbased AOPs.
scale-up. It is hard to find any studies which address both
economic feasibility and technical vitality of the UV-based
AOPs. An economically feasible process is one of the most
important aspects of a treatment system to be adopted in
industrial environment. The cost of treatment can be represented by the sum of the capital, operational, and maintenance costs. Figure 2 illustrates the example of a few
important components that are involved in cost estimation based on capital, operational, and maintenance cost
for UV-based AOPs. The cost of the processes used on the
UV system depends on the type of contaminants, properties of wastewater, flow rate of the effluents, and also the
design of the reactor (Rodrigues et al. 2014).
The costs can be calculated based upon the k
­ inetics
of degradation. From the literature studies, there are
two main orders proposed for the kinetics of AOPs:
pseudo-first-order and zero-order kinetics (Beak et al.
2009). UV combined with AOPs is an energy-intensive
process and has significant contribution to the operating
cost of the process. The information on the energy consumption by the treatment system can be very useful for
researcher to build the AOP-based wastewater treatment
plant system.
The figures of merit for AOPs based on the electrical energy consumption (system use electric energy) or
area collector (solar energy system) were proposed by
the International Union of Pure and Applied Chemistry
(IUPAC) for the evaluation and comparison of wastewater
treatment system (Bolton et al. 2001). Two figures of merit
were proposed for electrical driven system. Figure of merit
is a numerical quantity used to measure the efficiency
of the system based on the more characteristic system.
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A number of researchers were attempting this technique
to calculate the energy consumption based on the possible factors that may affect the efficiency such as average
UV radiation over the time, the volume of the water to
be treated, and the flow rate of the system. According to
figures of merit, first-order kinetics is used for low concentration of pollutants and zero-order kinetics for the high
concentration of pollutant.
The figure of merit used for highly concentrated
wastewater is electrical energy per mass (EEM). Zeroorder kinetic is applicable here, since the rate of removal
is directly proportional to the rate of electrical energy
consumption. EEM is defined by the amount of electrical
energy in kilowatt hours (kW h) required to degrade 1 kg
of the pollutant (Bolton et al. 2001). The simple formulas
to calculate EEM (kW h/kg) are
EEM =
EEM =
P × t × 1000
Batch scale
V × M ×( Ci -C f )
(46)
P × t × 1000
Continuous operation
V × F ×( Ci -C f )
(47)
Besides that, EEM is inversely proportional to factors
such as the photon flow, the fraction of light absorbed by
the reactor, and the quantum yield of the generated intermediates. The maximum efficiency and minimum feasible
value of EEM could be achieved when the system has a
larger amount of quantum yield and larger amount of light
absorbed in the system according to Eq. (48).
EEM =
P × 1000
M × 3600 ×( Gx-Z ) (48)
Electrical energy per order (EE/O) is defined as electrical energy (kW h) required to reduce the concentration of
a pollutant by one order of magnitude in 1 m3 (1000 ml)
of contaminated water (Bolton et al. 2001). This figure of
merit is best applied for the pollutant with low concentration. The high value of EE/O resembles a low energy efficiency of a system. This figure of merit is assumed to be
first-order kinetic as Eq. (18), t (min) is the reaction time in
the reactor, and k1′ is the first-order rate constant (min-1).
EEO =
EEO =
P × t × 1000
Batch scale
V × ln×( Ci -C f )
P × t × 1000
Continuous operation
F × ln×( Ci -C f )
ln Ci /C f = k1't (49)
(50)
(51)
where P is the rated power (kW), M is the molar mass
(g/mol), t is the irradiation time (min), V is the volume (l)
of the wastewater in the reactor, Ci and Cf are the initial
and final wastewater concentrations, and k1 is the pseudofirst-order rate constant (min-1) for the degradation of
wastewater.
It is obvious that figure of merit concept allows the
calculation of capital cost required by any UV-based AOP
system. This will be a great help for researchers to compare
the efficiency of the various systems based on the electrical
consumption to an individual system based on the order
of reaction (zero- or first-order kinetics). Although AOPs
are extensively studied for the degradation of a variety of
refractory wastewaters, only a number of studies focus on
the economic analysis. It is noted that most researchers
focus on the technical feasibility of the process by studying parameters optimization, better configuration of a
system, and design of a reactor, for a maximum efficiency.
A very prominent study was conducted to evaluate the
cost of UV-based AOPs based on the concept proposed by
IUPAC in this paper. As mentioned earlier, the economic
analysis is necessary to evaluate the system efficiency; a
good system should minimize the cost and maximize the
efficiency.
Mahamuni and Adewuyi (2010) have conducted a
very impressive review on the cost estimation for AOPbased ultrasound wastewater treatment. The authors have
reviewed the cost of the following UV system, although
the work was focused on the US for a comparison purpose:
(1) UV alone, (2) UV+US, and (3) US combined with UV
and O3. The costs have been calculated for flow rate of
1000 l/min. Rate constant was used as a basis to calculate
the cost. Time taken for the 90% degradation was considered as a residence time. The amount of energy required
to achieve 90% removal was calculated from the energy
density (W/ml) in this study. The authors have calculated
the cost of the treatment by taking into consideration the
capital cost and operating cost involved. There are few
conclusions that have been made based on the type of
pollutants and UV+US studies in this work. From the cost
estimation analysis provided in Table 8, it can be summarized that US combined with UV and H2O2 is more efficient and more economical compare to other combined
systems. The higher cost of US compare to UV is observed
in this study due to higher electrical energy and capital
cost required by the system for the treatment. So as a conclusion, energy consumption by a particular system is one
of the major factors contributing to the increase in cost of
treatment system besides other operating costs such as
chemical and maintenance costs.
Lucas et al. (2010) have conducted an estimation of
the operating costs of the, O3, O3/UV, and O3/UV/H2O2
processes based on the experiments carried out in the
pilot-scale reactor. The authors excluded maintenance,
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293
Table 8: Cost estimation of various type of UV-Based AOPs (Mahamuni and Adewuyi 2010).
Method
K (min-1)
Pollutant
UV
0.0021
0.0589
0.0124
0.0524
0.4418
0.02064
0.0869
0.0055
0.005
0.181
0.02171
0.1793
0.0207
0.433
0.0357
0.712
0.0757
Phenol
TCE
Dye
Phenol
TCE
Dye
Phenol
Dye
Phenol
TCE
Dye
Phenol
Dye
Phenol
Dye
Dye
Phenol
UV+H2O2
UV+O3
UV+US
US+UV+O3
Photocatalysis
UV+US+H2O2
US+photocatalysis
Capital cost
($/1000 gallons)
Operating and maintenance
($/1000 gallons)
Total cost
($/1000 gallons)
5.48E+08
5.21E+06
9.09E+07
2.93E+07
6.96E+05
1.12E+07
1.35E+07
5.86E+09
3.43E+09
2.99E+07
1.50E+09
9.68E+07
2.67E+07
3.12E+09
7.99E+07
1.11E+08
4.55E+09
1.67E+08
2.29E+05
3.05E+06
4.05E+07
4.06E+05
3.57E+06
4.25E+06
1.73E+08
1.58E+08
1.18E+06
4.43E+07
4.65E+06
8.14E+07
9.51E+08
2.62E+06
2.36E+07
6.54E+08
7.15E+08
5.44E+06
9.40E+07
6.98E+07
1.10E+06
1.48E+07
1.78E+07
6.03E+09
3.59E+09
3.11E+07
1.54E+09
1.01E+08
1.08E+08
4.07E+09
8.25E+07
1.35E+08
5.20E+09
TCE, trichloroethylene.
capital, labor, and depreciation costs in their study. The
operational cost is evaluated based on the costs for electricity, oxygen production, lamp replacement, and H2O2
to operate the pilot plant. Based on the economic analysis of the investigated AOPs in this study, it revealed the
O3/UV/H2O2 to be the most economical process (1.31 euro
m-3 g-1 of TOC mineralized at pH 4 and a COD/H2O2 ratio
of 2). The operation cost was calculated based on the
TOC removal efficiency of different process applied in
this investigation; O3/UV/H2O2 was found to be the most
efficient, and although it used more energy compared to
O3 or O3/UV, O3/UV/H2O2 is still considered as the most
economical in this study.
Durán et al. (2012) have evaluated the costs of treating real effluents from an integrated gasification combined cycle power station. The operational costs based
on the consumption of electrical energy, reagents, and
catalysts were calculated from the optimal conditions of
each process studied. The authors have studied the following process and estimated the cost of treatment: (i)
a photo-Fenton process at an artificial UV pilot plant,
(ii) a modified photo-Fenton process with continuous
addition of H2O2 and O2to the system, and (iii) a ferrioxalate-assisted solar photo-Fenton process at a compound parabolic collector pilot plant. The economic
analysis was carried out by analyzing the degradation of
TOC present in wastewater. The operational costs were
compared with the amount of TOC removed in grams.
Based on this study, it was found that the cost increases
with the amount of TOC removed due to the energy
consumption by an artificial UV lamp. Although modified photo-­Fenton process is capable of reaching higher
mineralization degrees over a shorter period of time,
solar photo-Fenton/ferrioxalate-assisted process was
found to be a more profitable system with the treatment
cost of 6 €/m3 (for 75% mineralization).
The cost estimations of various AOPs in combination
with UV is crucial for maximizing the degradation efficiency and minimizing the overall cost of the treatment
system. In order to evaluate the efficiency of the process,
more pilot-scale studies need to be carried out using different processes such as UV-Fenton, UV/H2O2/O3, and UV/TiO2.
This is important for an industrial scale-up of the system.
5 Conclusions
This review constitutes reference documents in the field
of UV coupled with AOPs to help researchers develop or
design new technology utilizing both UV and AOPs to
treat recalcitrant wastewaters. UV-based AOPs have been
proven to be an efficient and a sustainable alternative for
degradation of recalcitrant contaminates compared to
the use of UV alone. But it should be noted that very few
studies have been conducted to evaluate the economical
feasibility of this imperative technology.
The following are the main conclusions of this study:
1. UV/H2O2 is efficient because of rapid production of
hydroxyl radical in the reaction medium. However, its
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A. Buthiyappan et al.: Degradation and cost of UV IAOP for WT
performance is strongly affected by experimental conditions such as UV light sources, pH, and the properties of targeted wastewater.
2. The UV-TiO2 process has proven to be an effective oxidation technique, but its application is limited and
not effective for highly concentrated pollutants. Major
drawbacks identified are low efficiency at high catalyst concentrations and complex separation and recycling of TiO2.
3. The combination of UV radiation with ozone is more
efficient than using ozone alone for certain recalcitrant pollutants due to the formation of additional
hydroxyl radical via photolysis. However, this method
is not economically viable because of the high energy
requirement for both ozone formation and UV light.
4. Utilizing UV light simultaneously with Fenton reagents is economical, viable, and efficient. This is
because Fenton reagents are readily available, safe to
handle, and non-toxic. Moreover, use of UV light can
possibly reduce the consumption of H2O2 in comparison with conventional Fenton oxidation. However,
in comparison with the TiO2/UV system, separation
of catalyst is easier. But limitations linked with conventional Fenton oxidation such as the initial pH of
the solution, scavenging of hydroxyl radicals by nontarget substances, and sludge formation still needs to
be considered for future applications.
5. Sulfate radicals also offer several advantages over
other oxidants such as longer half-life, fast ­kinetics,
higher stability than hydroxyl radical, greater transport distances in the sub-surface level, ability to
work in a wide range of pH, and ability to be activated by low-cost oxidant precursors. However,
although it is more stable than hydroxyl radical, the
narrow selectivity of sulfate radical toward organic
matter makes it less efficient compared to hydroxyl
radical. However, UV/S2O82- process is very sensitive
to the operating parameters; therefore, it is necessary to optimize the parameters to enhance the
removal efficiency. Besides, due to the complexity
of the reaction involved in the UV/S2O82- process,
researchers should focus more on the kinetics
involved in the reactions.
6. Recently, researchers were actively involved to
improve the efficiency of UV-based AOPs by combining with more than one energy-dissipating component (microwave, solar, and US), oxidants (H2O2 and
O3), and catalyst (TiO2, Fe(II), Fe(III), and zero iron).
This combined process shows better efficiency compared to a single system by maximizing the efficiency
of degradation and minimizing the cost. More studies
7.
need to be carried out to evaluate the reliability and
application of this process.
Additionally, the cost estimate of various AOPs in
combination with UV is crucial for maximizing the
degradation efficiency and minimizing the overall
cost of the treatment system. The industrial scale-up
of UV-based AOPs is complicated without sufficient
studies on cost evaluation, so more pilot studies
need to be carried out on economical feasibility of
UV-AOPs.
6 Recommendations
The foregoing discussion on wastewater treatment using
UV-based AOPs concludes that its application for real
wastewater treatment is challenging. There are a few limiting issues that are required to be explored for its largescale applications. Therefore, based on the literature
reviewed in this study, the following recommendations
should be considered for improving UV-AOP-based wastewater treatment:
1. Wastewater stream turbidity is one of the major
factors that have not been elucidated much in the
literature. The treatment system is less effective if
other organic matter is predominantly present. This
is because oxidant and catalyst requirement can be
exceedingly high in order to achieve effective degradation of trace target pollutants. Higher turbidity is caused by the presence of particulate matter
which results in less penetration of UV light through
wastewater. As a consequence, process efficiency is
dramatically decreased due to the less interaction
of UV light with oxidants and catalyst. In addition,
turbidity is also directly proportional to the concentration of the pollutant, so careful considerations
are required to optimize the parameters such as dosage of catalyst and oxidant as well as intensity of
UV radiation. Besides, turbidity also increases the
energy requirements of the process, as higher/strong
UV irradiation is required to sustain the overall efficiency of the process. Therefore, in order to increase
the effectiveness of the process, it is recommended
to use small volumes of the effluent streams. This
requires a careful consideration during the design
of treatment system, for example, use of stacked
tube reactors with multiple UV lights at different
locations.
2. Besides, it is important to study the kinetics of combined or hybrid UV-AOPs. CFD models can be used
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to determine the performance of different UV-AOP
systems. For scale-up, a proper and accurate kinetic
model needs to be developed based on the types of
pollutant as the reaction pathways solely depend on
the parental compounds, intermediates, and final
products. The intermediates formed during the reaction should also be analyzed. It is important because
they can increase the toxicity of the reaction solution
sometimes. Numerous analytical methods such as
GC, GC/MS, and FTIR can be used to determine the
organic or inorganic compounds.
3. A proper study on the scavenging effect is also suggested as it could adversely affect the treatment efficiency and cause the formation of by-products which
could possibly increase the COD, TOC, total suspended solids (TSS), and toxicity value.
4. Use of various UV-based technologies for the degradation of emerging pollutants such as alachlor, atrazine,
carbamazepine, sulfamethoxazole, and others should
also be conducted with possible aid from CFD.
5. The ·OH exposure distribution should be evaluated
for each UV-based system.
6. Since the mechanism of a combined system is very
complex and dependent on the time and other operating parameters, it is very important to develop a reliable model that is capable of optimizing the system’s
parameters. Therefore, more research should be carried out to design and develop comprehensive models
that can accurately predict the performance of a pilot
and full-scale system.
7. High water consumptions in industry cause the excessive and irresponsible use of ground water, so zero
water discharge would contribute to the conservation
and replenishment of groundwater resources. Since
in the zero water discharge system no effluent will be
produced, it could eliminate the cost required to process the discharge water after treatment. Since most
of the UV-AOPs system managed to achieve higher
COD and TOC value, it is suggested to look into the
modification of UV-based AOPs system by integrating
with filtration system and designing a hybrid system
which could give no discharge of water as the treated
water can be recycled back to the system.
Acknowledgments: The authors are grateful to the
University of Malaya High Impact Research Grant
(HIR-MOHE-D000037-16001) from the Ministry of Higher
Education Malaysia and University of Malaya Postgraduate Research Fund which financially supported this
work.
295
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Bionotes
Archina Buthiyappan
Faculty of Engineering, Department of
Chemical Engineering, University of Malaya,
50603 Kuala Lumpur, Malaysia
Archina Buthiyappan graduated with a Bachelor’s degree in Industrial Chemistry in 2008 and a Master’s degree in Forensic Science
in 2010 from University Technology of Malaysia. She joined the
University of Malaya, Malaysia, as a doctoral candidate in 2012. Her
research focuses include application of various types of advanced
oxidation processes such as Fenton, photo-Fenton, and electroFenton to treat real textile effluents.
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Abdul Raman Abdul Aziz
Faculty of Engineering, Department of
Chemical Engineering, University of Malaya,
50603 Kuala Lumpur, Malaysia,
azizraman@um.edu.my
Abdul Raman Abdul Aziz completed his PhD in the area of threephase mixing. Currently, he is a Professor and holds the position
of Deputy Dean at the Faculty of Engineering, University of Malaya,
Malaysia. His research interests are in advanced wastewater treatment and mixing in stirred vessels. Prior to joining UM, he worked
in the oil and gas and food industries from 1989 to 1993. He is also
active in consultancy projects and is currently supervising many
PhD candidates. He has to date published more than 100 papers in
journals and conference proceedings both locally and internationally. He is also a member of professional and learned societies
such as the Institution of Chemical Engineers (IChemE, UK) and the
Institution of Engineers Malaysia (IEM).
Wan Mohd Ashri Wan Daud
Faculty of Engineering, Department of
Chemical Engineering, University of Malaya,
50603 Kuala Lumpur, Malaysia
Wan Mohd Ashri Bin Wan Daud is a Professor of Chemical Engineering, University of Malaya, Malaysia. He obtained his Bachelor’s
degree in Chemical Engineering in 1991 from Leeds University,
Leeds, UK, and his Master’s degree in Chemical Engineering in
1992 from the University of Sheffield, Sheffield, UK. He earned his
PhD degree in Chemical Engineering in 1996 at the University of
Sheffield. His research fields include fuel cell, energy, biomass conversion and the synthesis of catalyst materials, catalysis, zeolites,
polymerization process, separation process (adsorption, activated
carbon, and carbon molecular sieve), ordered mesoporous materials, and hydrogen storage materials. Professor Daud has published
approximately 90 research papers.
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