17
CHAPTER 2
LITERATURE REVIEW
2.1 INTRODUCTION
The presence of organic contaminants in surface and ground water
may be due to contaminated soil, agricultural runoff, industrial wastewater
and hazardous compounds storage leakage. The presence of these organic
compounds in water poses serious threat to the public health since most of
them are toxic, endocrine disrupting, mutagenic or potentially carcinogenic.
Many organic pollutants are toxic and detrimental even when they are present
at very less concentrations. For this reason, their removal from the
contaminated water is of high priority. Consequently, the need for efficient
treatment of these contaminants is imperative. In certain cases, conventional
treatment methods such as biological processes are not effective due
to the recalcitrant nature of the contaminants present (Garg et al 2010;
Bernal-Martinez et al 2010). Therefore, oxidation processes are preferred to
degrade such organics present.
2.2 DIRECT OXIDATION PROCESSES
Direct oxidation processes are widely used to degrade
bio-refractory substances like phenolic compounds. The process conditions
and the limitations of direct oxidation methods for phenolic wastewater
treatment are summarized in Table 2.1. It can be inferred from the table that
high degradation efficiencies are possible with direct oxidation techniques.
However, pollution load, process limitations and operating conditions are the
18
key factors to be considered during the selection of most appropriate
oxidation process for a particular compound degradation. Apart from high
degradation efficiency, direct oxidation processes demand specified operating
conditions to degrade the target compounds and this will increase the
operating cost of the process.
Table 2.1 Process conditions and limitations of direct oxidation
methods employed for phenolic wastewater treatment
Process Conditions Efficiency Process limitations Reference
Incineration1000 –
1700°C,1 bar
>99%
Practically feasible if CODfeed >300 g/L
Needs external fuel source for dilute wastes
Generate more toxic chemicals
Cybulski(2007)
SupercriticalWet Air
Oxidation(SWAO)
400-650°C,>250bar
>99.9%
CODfeed>50,000 mg/LScaling problemCorrosive reaction
environmentN-containing organics
difficult to oxidize
Kritzer andDinjus(2001)
Wet AirOxidation(WAO)
150-350°C,20-200 bar
75-90%
500<CODfeed<15,000 mg/LHigh energy consumptionTotal mineralization isdifficult
Kolackowskiet al (1999)
CatalyticWet Air
Oxidation(CWAO)
120-250°C5-50 bar
75-99%CODfeed>10,000 mg/LCatalyst dependent processCatalyst instability problem
Levec andPintar(2007)
O3, H2O2 &Fenton
Ambientconditions
Up to100%
Use expensive chemicals such as H2O2, O3
Needs posterior separationof Fe from the effluentLow pH requirement
Malato et al(2002)
19
Advanced oxidation processes (AOP’s) are alternative wastewater
treatment processes, which are cabable to degrade biorefractory organic
compounds. AOP’s typically operate with less energy requirement than direct
oxidation.
2.3 ADVANCED OXIDATION PROCESSES (AOPs)
The AOP’s are usually conducted at near ambient temperature and
pressure which involves the generation of hydroxyl radicals in sufficient
quantity to oxidize pollutants present in wastewater (Glaze et al 1987). The
hydroxyl radicals are extraordinarily reactive species, which attack the organic
molecules with rate constants usually in the order of 106 – 109 M-1S-1 (Hoigne
1997). The versatility of AOP is also enhanced by the fact that they offer
different possible ways for hydroxyl radicals production and allowing a better
compliance with the specific treatment requirements. The reduction potential
of various oxidants is presented in Table 2.2. Hydroxyl radical is the second
strongest oxidant preceded by the fluorine, and it reacts 106-1012 times faster
than ozone depending on the substrate to be degraded (Munter 2001).
Table 2.2 Standard reduction potential of common oxidants (Weast 1977)
Oxidant Oxidation Potential (v)Fluorine(F2) 3.03Hydroxyl radical( OH) 2.80Atomic Oxygen(O) 2.42Ozone (O3) 2.07Hydrogen Peroxide(H2O2) 1.77Hydroperoxyle radical( O2H) 1.70(a)
Potassium permanganet(KMnO4) 1.67Chlorine dioxide(ClO2) 1.50Hypochlorous acid(HclO) 1.49Chlorine (Cl2) 1.36Bromin(Br2) 1.09 (a) Legrini et al 1993
20
The mechanism of oxidation reaction involving hydroxyl radical
and organic substrates in aqueous solutions may be classified into
three different categories (Buxton et al 1988; Bossmann et al 1998;
Pignatello et al 2006).
Abstraction of hydrogen atom from C-H, N-H or O-H bonds of
organic molecules;
ROHRHOH 2 (2.1)
Electron transfer to hydroxyl radicals;
OHRXRXOH (2.2)
Addition of one atom with multiple atom compound and this
often happens with aromatic structure or olefins;
reactionsfurthurHOPhPhOH (2.3)
After the initiation of hydroxyl radical reaction with organic
compounds, a series of degradation steps are followed to mineralize the
organics completely. Mineralization refers to the conversion of organics into
CO2 and water and the degradation mechanism of many organic compounds is
still unclear. The application of AOP’s depends on the polluting load of
wastes as shown in Figure 2.1
21
Figure 2.1 Suitability of water treatment technologies according to
COD contents (Andreozzi et al 1999)
The wastes with relatively less chemical oxygen demand (COD)
content ( 10.0 g/L) can be suitably treated by means of AOP techniques
(Andreozzi et al 1999) since higher concentration of COD would require the
consumption of too large amounts of expensive reactants. Wastewater
containing massive pollutants load can be more conveniently treated by wet
oxidation or incineration techniques.
Generally, AOPs are classified according to the reactive phase
(homogeneous and heterogeneous) or hydroxyl radical generation methods
(chemical, electro-chemical, sono-chemical and photochemical). The
classification of advanced oxidation processes based on the source used for
the generation of hydroxyl radicals is presented in Table 2.3. The
processes involving combined advanced oxidation processes like
photo-electro-Fenton and sono-electro-Fenton are also presented in Table 2.3.
The other AOPs not presented in the table include ionizing radiation,
microwaves and pulsed plasma techniques (Klavarioti et al 2009). In addition,
solar-irradiated processes were studied in order to decrease
the costs associated with the use of light from non-natural sources
22
(Anderson et al 1991). However, the solar energy based processes have
restricted applications in countries receiving less solar radiation.
Table 2.3 Classification of advanced oxidation processes (AOPs)
Type of process Example
Homogeneous
Fenton based processes
Fenton: H2O2+Fe2+
Fenton like: H2O2 +Fe3+/mn+
Sono-Fenton: US/H2O2+Fe2+
Photo-Fenton: UV/ H2O2+Fe2+
Electro-FentonSono-electro-FentonPhoto-electro-FentonSono-photo-Fenton
O3 based ProcessesO3
O3+UVO3+ H2O2
O3+ UV+H2O2
Heterogeneous
H2O2 +Fe2+/ Fe3+/mn+-solidTiO2/ZnO/Cds+UVH2O2 +Fe0/Fe(Nano zero valent iron)H2O2+ Immobilized nano zero valentiron
2.4 FENTON PROCESS
Fenton and related reactions encompass reactions of peroxides
(usually H2O2) with iron ions to form active oxygen species that oxidize
organic or inorganic compounds. The Fenton reaction was discovered by
H.J.H. Fenton in 1894 and he reported that H2O2 could be activated by
23
Fe2+ salts to oxidize tartaric acid (Fenton 1894). Forty years later, Haber and
Weiss discovered the mechanism of Fenton reaction and concluded that the
active oxidant generated by the Fenton reaction is the hydroxyl radical ( OH)
(Haber and Weiss 1934). Later Barb and his collaborators (Barb et al 1949,
1951a, 1951b) modified the reaction mechanism proposed by Haber
and Weiss and identified the active reagent as free radical. In the last few
decades, the importance of hydroxyl radical reactions has been recognized
and over 1700 rate constants for hydroxyl radical reactions with organic and
inorganic compounds in aqueous solution have been tabulated (Buxton et al
1988). In the recent past, Fenton reaction was efficiently utilized in
wastewater treatment process for the removal of many hazardous organics
from wastewater (Neyens and Baeyens 2003; Bautista et al 2008).
The traditionally accepted Fenton mechanism is represented in
Equations (2.4) to (2.12) and its reaction rates were well reported in the
literature (Sychev and Isak 1995). Equation (2.4) is recognized as Fenton
reaction and implies the oxidation of ferrous to ferric ions to decompose H2O2
into hydroxyl radicals. It is usually considered as the core reaction of the
Fenton chemistry. Furthermore, other reactions must be considered to
understand the whole process.
114.2
322
2 .8040, SmolLkOHOHFeOHFe (2.4)
The generated ferric ions can be reduced by reaction with excesshydrogen peroxide to form ferrous ion, and more radicals are produced asshown in Equation (2.5). This reaction is called Fenton-like reaction and isslower than the Fenton reaction, and allows Fe2+ regeneration in an effectivecyclic mechanism. In Fenton like reaction, apart from ferrous ionregeneration, hydroperoxyl radicals (•O2H) are produced. The hydroperoxylradicals may also attack organic contaminants, but they are less sensitive than
24
hydroxyl radicals. It should be noted that, the iron added in small amount actsas a catalyst while H2O2 is continuously consumed to produce hydroxylradicals.
).(1021, 1125.22
222
3 SmolLkHHOFeOHFe (2.5)
The following reactions are involved in Fenton chemistry.
).(108.56.2, 1186.2
32 SmolLkOHFeOHFe (2.6)
).(105.172.0, 1167.222
32
2 SmolLkOHFeHHOFe (2.7)
).(101.233.0, 1168.22
22
3 SmolLkHOFeHOFe (2.8)
Equations (2.5) to (2.8) represent the rate limiting steps in theFenton chemistry since they are responsible for the regeneration of ferrousions from previously produced ferric ions. Equations (2.9) to (2.12) alsoreported to occur during the Fenton process and they are radical–radicalreactions or hydrogen peroxide –radical reaction.
).(1085, 1199.222 SmolLkOHOHOH (2.9)
).(105.47.1, 11710.22222 SmolLkOHHOOHOH (2.10)
).(102.28.0, 11611.222222 SmolLkOOHHOHO (2.11)
).(104.1, 111012.2222 SmolLkOOHHOOH (2.12)
In the absence or presence of any organic molecule to be oxidized,the decomposition of hydrogen peroxide to molecular oxygen and wateroccurs according to Equation (2.13). This reaction leads to exploitation
25
of bulk oxidant and thus an unnecessary increase on treatment cost(Pignatello et al 2006).
OHOOH 2222 22 (2.13)
Equations (2.4) to (2.12) demonstrate that the Fenton process
follows a complex mechanism. The production of desired hydroxyl radical
occurs through the chain initiation reaction (Equation (2.4)). However
hydroxyl radicals can be scavenged by ferrous ions (Equation (2.6)),
hydrogen peroxide (Equation (2.10)), hydroperoxyl radicals (Equation (2.12),
and /or even may be auto scavenged (Equation (2.9)). The foregoing analysis
indicates that hydrogen peroxide may act both as radical generator
(Equation (2.4)) and as scavenger (Equation (2.10)).
Hydroxyl radicals may attack organic radicals produced from
organics present in the wastewater (Equations (2.1) to (2.3)). Those radicals
form dimmers or react with ferrous and ferric ions, as shown in Equations
(2.14) to (2.16) (Neyans and Baeyens 2003; Pignatello et al 2006).
RRR2 (2.14)
32 FeRFeR (2.15)
23 FeRFeR (2.16)
2.4.1 Stochiometric Relationship
The key features of the Fenton system are believed to be its reagent
conditions, i.e. [Fe2+], [Fe3+], [H2O2] and the reaction characteristics (pH,
temperature and the quantity of organic constituents). It is important to
understand the mutual relationship between these parameters in terms of
hydroxyl radical production and consumption. Yoon et al (2001) studied these
relationship and classified them into three categories according to the ratio
26
between initial concentration of Fe2+ and initial concentration of H2O2
([Fe2+]0/[H2O2]0). The results are summarized below.
2.4.1.1 [Fe2+]0/[H2O2]0 2
The Fenton reactions start with hydroxyl radicals production from
the reaction between ferrous ion and hydrogen peroxide (Equation (2.4)).
When the Fenton reaction initiated in the absence of organics under
[Fe2+]0/ [H2O2]0 2, the consumption ratio of ferrous ion to hydrogen peroxide
becomes about 2, and radical chain reactions are quickly terminated. This is
because the hydroxyl radicals produced as a result of Equation (2.4) mainly
react with the ferrous ion (Equation (2.6)) instead of hydrogen peroxide
(Equation (2.10)). This explanation is supported by the fact that the reaction
between hydroxyl radical and the ferrous ion is ten times faster than that
between hydroxyl radicals and hydrogen peroxide [k2.6 = 2.6 - 5.8.108 (L.mol-
1S-1), k2.10 = 1.7 - 4.5.107(L.mol-1S-1)].
On the other hand, the presence of organics affects the behavior of
ferrous ion whereas it does not affect the hydrogen peroxide. This is due to
the competence of organics towards hydroxyl radicals (Equation (2.1)). The
presence of organics reduces the consumption ratio to less than two
(d[Fe2+] / d[H2O2]) 1.3. This indicates that the ferrous ion is utilized as a
major reactant, not as a catalyst in the Fenton reaction.
Yoon et al (2001) investigated the effect of the Fenton reaction on
organics removal from landfill leachate. They have used a [Fe2+]0 / [H2O2]0
ratio of 1.25. In this case, the Fenton reaction was divided into two processes.
The first process is an initial oxidation at a pH of about 3. The second
process, which follows the oxidation process, is a coagulation, which occurs
at a pH of 7~8. It was interpreted that the coagulation step in the Fenton
27
reaction had a primary role in the selective removal of organics, even though
the Fenton reaction is not a coagulation process.
2.4.1.2 [Fe2+]0/[H2O2]0= 1
Regardless of the presence of organics, hydrogen peroxide rapidly
converts all ferrous to ferric ions (Equation (2.4)). In the absence of organics,
hydrogen peroxide decomposes slowly through ferric ion induced radical
chain reaction (Equation (2.5)). These reactions (Equation (2.4) and (2.5))
occur immediately after the consumption of hydrogen peroxide. The reduction
of ferric ion (Equation (2.5)) is significantly lower than Fenton reaction
(Equation (2.4)) and is the rate determining step. To have a continued
decrease of hydrogen peroxide, ferrous ion must be formed by the reduction
of ferric ion. Then, the Fenton reaction can be characterized by two specific
systems: ferrous system and ferric system. The occurrence of the system
depends on the oxidation stage of the iron initially added or the major
oxidation state of the iron present.
The ferrous system refers to the primary reaction, which produces
hydroxyl radicals (Equation (2.4)). The ferric system refers to the case in
which the ferric ion induced Equation (2.5) must be preceded in order to
produce hydroxyl radicals via Equation (2.4).
However, the presence of organics has an impact on the behavior of
hydrogen peroxide in two ways: (i) no further hydrogen peroxide
decomposition occurs just after the initial decrease of hydrogen peroxide,
since the reaction of organics with hydroxyl radicals (Equation (2.1))
overwhelms the reaction of hydrogen peroxide with hydroxyl radicals
(Equation (2.10)); (ii) the presence of excess organics hinders the reaction
between hydroxyl radicals and ferrous ion. Therefore, the remaining ferrous
ion reacts with the hydrogen peroxide and shows a slightly higher
28
consumption of hydrogen peroxide at the initial stage of the reaction (Yoon et
al 2001).
2.4.1.3 [Fe2+]0/[H2O2]0<< 1
In the absence of organics, a slow decomposition of hydrogen
peroxide is caused by ferric ion, which induces radical chain reactions (ferric
system) immediately after the initial depletion of hydrogen peroxide.
However, the presence of organics stops the decomposition of hydrogen
peroxide by ferric ion. At a low consumption ratio of ferrous to hydrogen
peroxide, hydroxyl radicals react with hydrogen peroxide to a greater extent
producing hydroperoxide radicals via Equation (2.10). These hydroperoxide
radicals are further used to participate in propagating radical chain reaction by
reducing ferric to ferrous ion (Equation (2.8)), and results in a larger
consumption of hydrogen peroxide than that occurs in the absence of
organics. The studies reported in the literature with respect to the degradation
of phenolic effluents by Fenton process are summarized in Table 2.4.
29
30
31
32
33
2.4.2 Optimum Operating Conditions of Fenton Reaction
Based on the exhaustive analysis of the existing literature on the
application of Fenton oxidation to wastewater treatment, following optimum
conditions is outlined.
2.4.2.1 Operating pH
Fenton process is strongly dependent on the solution pH mainly due
to iron and hydrogen peroxide speciation factors. The optimum pH for the
Fenton reaction was found to be around 3, regardless of the target substrate
(Eisenhauer 1964; Ma et al 2000; Rivas et al 2001).
The activity of Fenton reagent is reduced at higher pH due to the
presence of relatively inactive iron oxohydroxides and formation of ferric
hydroxide precipitate (Parsons 2004). In this situation, less hydroxyl radicals
are generated due to the presence of less free iron ions. The oxidation
potential of hydroxyl radicals decreases with increasing pH. The oxidation
potential for the redox couple OH/ H2O has been reported to be 2.59 V at
pH 0 and 1.64 V at pH 14 (Bossmann et al 1998). In addition, auto-
decomposition of hydrogen peroxide is accelerated (Equation (2.13)) at high
pH (Szpyrkowicz et al 2001). At pH below 3, decrease in degradation
efficiency was observed (Kavitha and Palinivelu 2005). At very low pH
values, iron complex species [Fe (H2O) 6]2+ exist, which reacts more slowly
with hydrogen peroxide than other species. This phenomenon was also
influenced by the concentration of ferrous ion present. In addition , the
peroxide gets solvated in the presence of high concentration of H+ ions to
form stable oxonium ion [H3O2]+. Oxonium ions make hydrogen peroxide
more stable and reduce its reactivity with ferrous ions (Xu et al 2009; Kwon
et al 1999). Therefore, the efficiency of the Fenton process to degrade organic
compounds is reduced both at high and low pH. Thus an adequate control of
34
pH would increase process efficiency. However, reaction buffering will
increase the operating costs. Therefore, final decision of utilization of buffers
varies with situation.
Fenton process can be carried out at room temperature and
atmospheric pressure. In addition, required reagents are readily available, easy
to store and handle, safe and they do not cause environmental damages
(Pignatello et al 2006). However, two main draw backs were identified. The
first is related to the wastage of oxidants due to the radical scavenging effect
of hydrogen peroxide (Equation (2.10)) and its self decomposition (Equation
(2.13)). The second refers to the continuous loss of iron ions and the
formation of solid sludge. Several economic and environmental draw backs
have been reported to occur with Fenton sludge (Benatti et al 2009). Thus,
technologies allowing an efficient use of H2O2 have to be studied.
Furthermore, an attempt has to be made for the recovery of iron ions and their
subsequent recycle and reuse.
2.4.2.2 Amount of ferrous ions
Usually the rate of degradation increases with an increase in the
concentration of ferrous ion (Lin and Lo 1997). However, the extent of
increase is sometimes observed to be marginal above a certain concentration
of ferrous ion as reported by Lin et al (1999), Kang and Hwang (2000) and
Rivas et al (2001). Also, an enormous increase in the ferrous ions will lead to
an increase in the unutilized quantity of iron salts, which will contribute to an
increase in the total dissolved solids content of the effluent stream and this is
not permitted. Thus, laboratory scale studies are required to establish the
optimum loading of ferrous ions to mineralize the organics.
35
2.4.2.3 Hydrogen peroxide concentration
Concentration of hydrogen peroxide plays a more crucial role in
deciding the overall efficiency of the degradation process. Usually it has been
observed that the percentage degradation of the pollutant increases with an
increase in the dosage of hydrogen peroxide (Lin and Lo 1997; Lin et al 1999;
Kang and Hwang 2000; Rivas et al 2001). However, care should be taken
while selecting the operating oxidant dosage. The residual hydrogen peroxide
contributes to COD (Lin and Lo 1997) and hence excess amount is not
recommended. Also, the presence of hydrogen peroxide is harmful to many of
the organisms (Ito et al 1998) and will affect the overall degradation
efficiency significantly, where Fenton oxidation is used as a pretreatment to
biological oxidation. Another negative effect of hydrogen peroxide is the
scavenging of generated hydroxyl radicals, which occurs at large quantities of
hydrogen peroxide. Thus, the dosage of hydrogen peroxide should be adjusted
in such a way that the entire amount is utilized and this can be decided based
on the laboratory scale studies.
2.4.2.4 Initial concentration of pollutant
Usually lower initial concentrations of the pollutants are favored
(Kwon et al 1999; Benitez et al 2001), but the negative effects of treating
large quantity of effluent needs to be analyzed before the dilution ratio is
fixed. For real industrial wastewater, dilution is essential before any
degradation is effected by Fenton oxidation.
2.4.2.5 Type of buffer used for pH adjustment
As discussed earlier, degradation rates are observed to be higher at
pH of approximately 3 and hence, operating pH should be maintained around
the optimum value. It should be noted that the type of buffer solution used
also has an effect on degradation process (Benitez et al 2001). The acetic
36
acid/acetate buffer gives maximum oxidation efficiency whereas least is
observed with phosphate and sulfate buffers (Benitez et al 2001). This can be
attributed to the formation of stable Fe3+ complexes that are formed under
those conditions (Pignatello 1992).
2.4.2.6 Operating temperature
Limited studies are available depicting the effect of temperature on
the degradation rates. Moreover, ambient conditions can safely be used with
good efficiency. In fact, Lin and Lo (1997) reported an optimum temperature
of 30 C, whereas Rivas et al (2001) reported that the degradation efficiency is
unaffected even when the temperature is increased from 10 to 40 C. If the
reaction temperature is expected to rise beyond 40 C due to exothermic
nature, cooling is recommended. The efficient utilization of hydrogen
peroxide decreases due to accelerated decomposition of hydrogen peroxide
into water and oxygen (Nesheiwat and Swanon 2000).
2.4.2.7 Chemical coagulation
Chemical coagulation step is recommended after the Fenton
oxidation so as to keep the concentration of the soluble iron with the specified
limits (Lin and Lo 1997). Lin et al (1999) have also demonstrated the efficacy
of chemical coagulation in controlling the concentration of total dissolved
solids below the specified limits.
2.5 SONO-FENTON PROCESS
The oxidation of organics by ultrasound has received considerable
attention because of its rapid degradation of chemical contaminants (Francony
and Petrier 1996). Ultrasound is a sound wave with a frequency greater than
the upper limit of human hearing (approximately 20 kHz). In practice, three
frequency ranges of ultrasound are reported for three different uses (Mason
37
and Cordemans 1996): (1) the relatively low frequency range , which is
applied for conventional power ultrasound (20-100 kHz), (2) the medium
frequency range, which is used for sonochemical effects (300-1000 kHz), (3)
and the high frequency range, which is typically used for diagnostic
imaging(2-10MHz).
The application of ultrasound wave creates expansion and
compression cycles. The expansion cycle causes reduction of pressure in the
liquid, and, if the amplitude of ultrasound pressure is sufficiently large, it can
result in acoustic cavitation, a process of formation, growth, and implosion of
bubbles filled with vapor and /or gas. The growth and implosion of bubbles
are affected by physical properties of gas and liquid, initial size of gaseous
nuclei present in liquid, ultrasound frequency and intensity. When these
cavitations bubbles explosively collapse, the pressure and temperature in the
bubbles reach up to several hundred atmosphere and several thousand Kelvin,
respectively (Susick et al 1990; Dahlem et al 1998). Under these conditions,
organic compounds are decomposed directly by pyrolytic cleavage. On the
other hand, the hydroxyl radicals formed by pyrolysis also help to degrade the
organics. Thus, in sonochemistry, there are three potential reaction sites: (1)
inside of cavitation bubbles, (2) interfacial region between the cavitation
bubble and liquid phase, and (3) bulk liquid (Liang et al 2007).
The hydroxyl radicals are generated by water pyrolysis as shown in
Equation (2.17). After that, the radicals can enter into a variety of chemical
reactions in a gas bubble and/or in the bulk solution. Some typical reactions
involved are shown in Equations (2.18)-(2.21) (Fisher et al 1986; Serpone
et al 1994).
38
HOHOH2 (2.17)
OHHOH 2 (2.18)
OOHOH 22 (2.19)
222 OHOH (2.20)
22 HH (2.21)
Hydrophobic chemicals with high vapour pressure have been
proved to undergo mainly thermal decomposition inside the cavitation
bubbles (Kotronarou et al 1991; Drijvers et al 1999). Whereas, hydrophilic
compounds with low vapour pressures tends to remain in the bulk solution.
Therefore, the major reaction site for these compounds is the bulk liquid
region, where they can be easily destroyed by an oxidative degradation. But,
sufficient quantities of hydroxyl radicals are to be ejected into the liquid phase
during the collapse of cavities.
To increase the hydroxyl radical concentration in the bulk solution,
Fenton and sonolysis can be combined together. These methods utilize the
advantages of ultrasound and Fenton’s reagent, allowing improved
degradation of organic pollutants (Liang et al 2007). The studies reported in
the literature related to sono-Fenton process on organic compounds
degradation are summarized in Table 2.5
39
40
41
42
43
2.6 ELECTRO-FENTON PROCESS
There is a greater interest in the development of effective
electrochemical treatments for the destruction of toxic and biorefractory
organics (Pulgarin and Kiwi 1996). Anodic oxidation and indirect electro-
oxidation are the most usual techniques utilized to achieve the mineralization
of such pollutants. In anodic oxidation, pollutants are mineralized by direct
electron transfer reactions or action of radical species (hydroxyl radicals)
formed on the electrode surface as shown in Equation (2.22).
eHOHOH ads2 (2.22)
This radical, as discussed previously, is a powerful oxidizing agent,
with an ability to react with organics giving dehydrogenated or hydroxylated
derivatives.
Anodic oxidation (Comninellis and Pulgarin 1991, 1993; Morselli
et al 2003; Pannizza and Cerisola 2005) is usually performed in the anodic
compartment of the divided cell, where the contaminated solution is treated
using an anode of Pt, undoped and doped PbO2 and SnO2. Under these
conditions, however, most aromatic solutions are poorly mineralized, because
of the generation of hardly oxidizable carboxylic acids.
In electro-Fenton process, pollutants are destroyed by the action of
Fenton’s reagent in the bulk together with anodic oxidation at the anode
surface. Electro-Fenton process is classified into four categories depending on
Fenton reagent’s addition or formation.
In type 1, hydrogen peroxide and ferrous ion are electro-generated
using a sacrificial anode and an oxygen sparging cathode, respectively (Ting
et al 2008).
44
In type 2, hydrogen peroxide is externally added while ferrous ion
is produced from sacrificial anode as shown in Equation (2.23) (Kurt et al
2007).
eFeFe 22 (2.23)
In type 3, ferrous ion is externally added and hydrogen peroxide is
generated using an oxygen sparging cathodes (Brillas and Casado 2002;
Badellino et al 2006).
In type 4, hydroxyl radical is produced using Fenton reagent in an
electrolytic cell and ferrous ion is regenerated through the reduction of ferric
ions on the cathode (Zhang et al 2006, 2007).
However, electro-Fenton processes have some problems related to
H2O2 production. The production of H2O2 is slow because oxygen has low
solubility in water and the current efficiency under reduced pH (pH<3) is low
(Ting et al 2008). The most promising method is type 4, in which ferric ion is
reduced to ferrous ion at the cathode; however, Fe2+ regeneration is slow even
at an optimal current density (Ting et al 2008). The efficiency of electro-
Fenton process depends on electrode nature, pH, catalyst concentration,
electrolytes, dissolved oxygen level, current density and temperature.
The studies related to electro-Fenton process reported in the
literature are summarized in Table 2.6.
45
46
47
48
49
50
2.7 PHOTO-FENTON PROCESS
A combination of hydrogen peroxide and UV radiation with Fe2+ or
Fe3+ oxalate ion (photo–Fenton process) produces more hydroxyl radicals
compared to conventional Fenton method or photolysis and in turn increases
the rate of degradation of organic pollutants (Ruppert et al 1993; Sun and
Pignatello 1993; Gogate and Pandit 2004). Fenton reaction accumulates Fe3+
ions in the system and the reaction does not proceed once all Fe2+ ions are
consumed. The photochemical regeneration of ferrous ions (Fe2+) by photo-
reduction (Equation (2.24)) of ferric ions (Fe3+) occurs in photo-Fenton
reaction (Faust and Hoigne 1990). The newly generated ferrous ions react
with H2O2 and generate hydroxyl radical and ferric ion, and the cycle
continues.
OHFehOHFe 22)( (2.24)
The studies reported in the literature showed that the combination
of Fenton reaction with UV irradiation gives a better degradation of organic
pollutants such as 4-chlorophenol (Bauer and Fallmann 1997), nitrobenzene
and anisole (Zepp et al 1992), herbicides (Sun and Pignatello 1993) and
ethyleneglycol (McGinnis et al 2000). Direct photolysis of H2O2 (Equation
(2.25)) produces hydroxyl radicals which can be used for the degradation of
organic compounds. However, in the presence of iron complexes, which
strongly absorb radiation, this reaction will contribute only to a lesser extent
for the photo-degradation of organic contaminants (Safarzadeh-Amiri et
al1997; De Oliveira et al 2007).
OHhOH 222 (2.25)
The photo-Fenton process offers better performance at pH 3.0, at
which the hydroxy-Fe3+ complexes are more soluble and Fe(OH)2+ are more
51
photoactive (Kim et al 1997; Legrini et al 1993). The photo-Fenton process
was reported as more efficient than Fenton treatment (Gogate and Pandit
2004). In some cases, use of sunlight instead of UV irradiation reduced the
costs. However, this offers a lower degradation rate of pollutants.
De Oliveira et al (2007) compared the performance of Fenton and
photo-Fenton processes for the treatment of painting industry effluent
(COD = 80.75 mg/L) and reported higher COD and TOC removal with solar-
assisted photo-Fenton process compared to Fenton treatment or when an
artificial radiation source was used. The performance comparison study made
with three iron sources such as FeSO4, Fe(NO3)3 and potassium ferrioxalate
(K3[Fe(C2O4)3], obtained by mixing Fe(NO3)3 with K2C2O4 solutions). The
formation of Fe3+ complex when Fe(NO3)3 was used resulted in poor
performance, whereas the addition of K3[Fe(C2O4)3] increased the carbon
loading of the wastewater. In the presence of 15 mM of Fe2+ and 300 mM of
H2O2, 99.5% COD reduction was reported when the wastewater irradiated
with solar radiation for 6 h.
Amat et al (2004) compared the degradation of two commercial
anionic surfactants such as sodium dodecyl sulphate and
dodecylbenzenesulphonate using Fenton reagents (Fe2+ or Fe3+ with H2O2 in
the presence or absence of solar radiation), photo-catalysis (TiO2 with solar
irradiation) and photo-degradation using solar sensitizer (pyrylium salt). They
demonstrated that the addition of the solar sensitizer did not efficiently
degrade the surfactants and their further studies concluded that the photo-
Fenton processes using solar radiation (0.1 mM of Fe2+ or Fe3+, and
1mM H2O2) had a higher rate of surfactant degradation than that of solar-TiO2
treatment. The illustrative works related to photo-Fenton process reported in
the literature are presented in Table 2.7.
52
53
54
55
2.8 SONO-PHOTO-FENTON PROCESS
The combined application of ultrasound and ultraviolet irradiation
along with Fenton reagent is known as sono-photo-Fenton process, which
enhanced the production of hydroxyl radicals in an aqueous system
significantly. Sonolysis of water produces hydroxyl radicals and hydrogen
atoms. However, significant loss of H and OH species occurs due to the
recombination. On the other hand, the application of UV irradiation,
converted the hydrogen peroxide produced by recombination of hydroxyl
radicals, and in turns increased the amount OH (Wu et al 2001). The
intermediate complex formed due to the reaction of Fe3+ with H2O2 during the
Fenton reaction could be reduced to Fe2+ by sonolysis (Pradhan and Gogate
2010) and photolysis (Ting et al 2008).
The direct photolysis ( <400 nm) of H2O2 also generates hydroxyl
radicals (Equation (2.25)) and the process is cyclic with respect to Fe2+ ions.
The amount of ferrous salt required under SPF condition is very small
compared to Fenton condition, where ferrous ions need to be added at regular
intervals to continue the reaction. Otherwise, reaction will stop after the
complete conversion of ferrous ions into ferric ions. SPF process reduces the
amount of ferrous ions present in the treated water, and this is very important
in an industrial point of view (Zepp et al 1992; Tao et al 2005; Collins 2002;
Chen et al 2002; Ma et al 2003; Bozzi et al 2004; Vaishnave et al 2012).
Segura et al (2009) investigated the application of SPF process for
the degradation of phenol using Fe2O3/SBA-15 as a heterogeneous catalyst.
The initial pH was maintained at 3. Total organic carbon (TOC) degradation
of 45% was observed with sono-photo-Fenton process whereas 30 and 40%
TOC degradation was observed with sono-Fenton and photo-Fenton processes
respectively.
56
Mendez-Arriaga et al (2009) investigated the degradation of
recalcitrant pharmaceutical micro-pollutant ibuprofen by means of sono-
photo-Fenton, sono-photo-catalysis and TiO2/Fe2+/sonolysis processes. The
presence of ultrasound irradiation in photo-Fenton process improved the iron
catalytic activity and ibuprofen degradation and mineralization to 95 and
60%, respectively. On the other hand, total removal of ibuprofen and
elimination of more than 50% of dissolved organic carbon were observed by
photocatalysis with TiO2 in the presence of ultrasound irradiation. The results
showed that, the hybrid system as a promising method for complete
elimination/mineralization of the recalcitrant micro-contaminant ibuprofen.
Vaishnave et al (2012) conducted a study on degradation of
azure-B by sono-photo-Fenton and photo-Fenton processes. The degradation
rate was found to strongly dependent on initial pH, initial concentration of
azure-B, Fe2+concentration, H2O2 concentration and UV light intensity. The
effect of these parameters was studied and the optimum operational
conditions of these processes were found. The optimum conditions obtained
for 1.33x10-4 mol/L of azure-B was: pH 2.1; H2O2 0.5 ml; Fe3+ 5.0 x 10-4 mol /L;
Ultrasonic frequency 40 kHz; and light intensity 75.5 mW/cm2. The results
showed that the dye was completely oxidized and degraded into CO2 and
H2O.
Zhong et al (2011) studied the degradation of C.I Acid Orange 7 by
the integrated heterogeneous sono-photo-Fenton process using mesostructured
silica (Fe2O3/SBA-15) as heterogeneous catalyst. The effect of hydrogen
peroxide concentration, initial pH, Fe2O3/SBA-15 loading, ultrasonic power
and initial solution concentration on the decolorization was investigated. The
results showed that the degradation efficiency increased with an increase in
hydrogen peroxide concentration, ultrasonic power, Fe2O3/SBA-15 loading,
but decreased with an increase in initial pH and initial dye concentration. The
57
particle size distribution of spent catalyst evidenced a remarkable size
reduction during reaction, leading to more reaction surface and higher mass
transfer rate. The reaction intermediates were identified using GC-MS and
degradation pathway was proposed.
2.9 SONO-ELECTRO-FENTON PROCESS
Many researchers have reported the coupling strategy between
sonochemistry and different AOPs. The combination of sonolysis and Fenton
process offered the concept of advanced sonochemical hybrid techniques that
possess significantly greater efficacy for water remediation (Joseph et al
2000; Minero et al 2005; Adewuyi 2005). A well known effect of ultrasound
applied to aqueous solution is cavitation. This phenomenon consists of the
formation, growth and collapse of micro bubbles that concentrate the acoustic
energy into the reactors. Hydroxyl radicals produced by water decomposition
(Equation (2.26)) are used for the degradation of organics (Adewuyi 2005).
HOHOH )))2 (2.26)
Abdesalam and Birkin (2002) studied the sono-electro-chemical
degradation of Meldola Blue using a recirculation type three electrode flow
cell with Pt gauze as the anode and reticulated vitreous carbon as the cathode
(15 mm x10 mm x5 mm). The reactor was coupled to a piezoelectric
transducer of 124 kHz and 100 V for treating 100 ml of 0.1 mM dye solution
with different Fe2+ concentration at pH 2 and 25°C. An apparent rate constant
of 2.4x10-2 min-1 was reported for the decolourization with 0.5 mM Fe2+
(optimum catalyst content) at Ecat=-0.7V/Ag, which is twice the value
obtained for Electro-Fenton process. Under these conditions, the COD was
reduced by 61% after 100 min of sono-electro- Fenton treatment.
58
Yasman et al (2004) studied the degradation of common herbicide,
2, 4-dichlorophenoxyacetic acid (2,4-D) and its derivative 2,4- dichloro
phenol (2,4-DCP) using SEF process. A cylindrical glass reactor containing
ultrasonic generator (75 W and 20 kHz) was immersed from the top of the
reactor. Both cathode and anode were made up of nickel foil (0.125 mm thin)
in the form of cylindrical segment of 11 mm radius and 20 mm height.
Temperature of the solution was maintained at 25°C by circulating water in a
double jacket cooling array. Complete degradation of 2, 4-D (1.2 mM) was
found using 3 mM Fe2+ and H2O2 at a reaction time of 60 sec. Thus, the time
required for the degradation was considerably shorter for SEF scheme when
compared to F and SF scheme. This result indicates that the SEF process is an
effective method for the detoxification of wastewater.
Oturan et al (2008) investigated the effect of EF and SEF process
for the degradation of herbicides namely 2, 4-dichlorophenoxyacetic acid
(2, 4-D) (1 mM)), 4, 6-dinitro-o-cresol (0.5 mM) and Azo benzene (0.1 mM).
The process was carried out using 0.1 mM Fe3+ in O2 saturated solution with
0.05M Na2SO4 at pH 3 and at 200 mA current intensity. For SEF process,
either a low frequency of 28 kHz or high frequency of 460 kHz was applied
with in an output power range from 20-80 kW. Complete degradation of
2, 4-D was achieved in 80 min using 20 W output. The results showed that the
nature of the organic structure plays an important role in SEF. The
improvement yielded by SEF at 20 W was ascribed to: i) the enhancement of
the transfer of reactants Fe3+ and O2 toward the cathode for the electro
generation of Fe2+and H2O2; ii) The additional generation of hydroxyl radicals
by sonolysis of water and iii) the pyrolysis of organics due to the caviataion
promoted by ultrasound irradiation.
Li et al (2010) evaluated the effect of low frequency ultrasonic
irradiation on the sono-electro-Fenton oxidation of cationic red X-GRL.
Ultrasonic irradiation significantly increased the hydrogen peroxide
59
production rate and reduced the time needed to reach the maximum hydrogen
peroxide concentration. In addition, ultrasonic irradiation has a considerable
effect on the degradation of cationic red X-GRL. The results showed that the
degradation rate followed pseudo-first order kinetics and also decolorization
rate increased with ultrasonic power. Furthermore, total organic carbon
removal efficiency and mineralization current efficiency were greatly
promoted in sono-electro-Fenton process compared to electro-Fenton process.
These results proved that sono-electro-Fenton process is a promising
technology in terms of colored wastewater treatment.
Martinez and Uribe (2012) investigated the degradation of azure B
dye by Fenton, sonolysis and sono-electro-Fenton process employing
ultrasound at 23 kHz and the electrolytic generation of hydrogen peroxide at
the reticulated vitreous carbon electrode. The dye degradation followed
apparent first-order kinetics in all the degradation processes. The rate constant
was affected by pH and initial concentration of ferrous ion, with the highest
degradation obtained at pH between 2.6 and 3. The rate constant for Azure B
degradation by sono-electro-Fenton is 10 times higher than that of sonolysis
and 2 times larger than that of Fenton process. The COD removal of 68 % and
85 % was obtained by Fenton and sono-electro-Fenton process respectively
and 90 % removal of Azure B was observed with both the processes.
2.10 PHOTO-ELECTRO-FENTON PROCESS
The catalytic effect of Fe2+ in the electro-Fenton process can be
enhanced by irradiating the contents with UV light. Thus, the combination of
electrochemical, photochemical and Fenton process, called photo-electro-
Fenton (PEF) process, generates greater quantity of free
radicals due to the synergistic effect (Boye et al 2002). The direct photolysis
of acid solution containing peroxide generates hydroxyl radicals through the
hemolytic breakdown of the peroxide molecule according to the
60
Equation (2.25). This reaction increased the oxidative capability of the
process due to the additional production of hydroxyl radicals. Thus, the
degradation of target organic substrate can be enhanced when the solution is
irradiated with UV light in addition to the application of electro-Fenton
process.
Photochemical regeneration of Fe2+ by the photo-reduction of Fe3+
ions and photo-activation of complexes renders the photo-electro-Fenton
systems more efficient (Brillas et al 2003a; Boye et al 2003). At acidic pH,
oxalic acids behave as the photo-active complexes in the presence of ferric
ions which undergo photo-decarboxylation reaction (Pignatello et al 2006) as
shown in Equation (2.27).
222 )()()()( CORIIFeCORhIIIFeCOR (2.27)
The studies pertaining to the application of photo-electro-Fenton
process are very limited and most of the studies are related to the treatment of
herbicide (Boye et al 2002, 2003; Irmak et al 2006; Brillas et al 2003b), and
dyes (Flox et al 2006). Flox et al (2007) recently used solar photo energy as
photon source and reduced the operating costs of the process substantially.
The various studies carried out using photo-electro-Fenton process are
summarized in Table 2.8. Some of the illustrative works in recent years have
been discussed in detail so as to establish the optimum operating parameters
and other critical considerations for attaining maximum efficiency of this
hybrid oxidation technique.
61
62
63
64
65
2.11 ZERO VALENT IRON MEDIATED DEGRADATION
Homogeneous oxidation processes were effectively used to degrade
wide variety of organics. But these processes are produce sludge in the final
neutralization step (Iurascu et al 2009). In order to avoid these disadvantages,
heterogeneous based AOPs could be used. In heterogeneous based AOPs,
nano zero valent iron (NZVI) was used to induce Fenton oxidation.
Zero valent state metals (such as Fe0, Zn0, Sn0 and Al0) are
surprisingly effective agents for the remediation of contaminated ground
water (Powell et al 1995; Warren et al 1995). In particular, it has been the
subject of numerous studies over the last 10 years and this is an interestingly
popular choice for treatment of hazardous and toxic wastes, and for the
remediation of contaminated sites.
2.11.1 Synthesis of Nano Zero Valent Iron (NZVI)
The earliest methods used for the production of well-defined iron
nano-particles relied upon dispersing them in mercury. A number of mercury
based methods were in use in the 1940s and 1950s (Huber 2005). However,
mercury-based methods have not been used because of the toxicity of
mercury vapors.
Recently researchers have developed different synthesis methods
that can be classified into two general categories, physical and chemical
synthesis method (Li et al 2006a). Such methods include, for example
precipitation from FeCl3 using sodium borohydride (Wang and Zhang 1997)
and high energy milling (Hakl et al 2002). Although, primary particle size,
particle size distribution, and degree of agglomeration/aggregation of NZVI
are different between these methods, each of these approaches has been
successfully applied to the preparation of NZVI particles.
66
2.11.2 Physical Synthesis Methods
2.11.2.1 High energy ball milling
High energy ball milling uses conventional mechanical grinding
techniques to break coarse metal grains into nano scale particles (Hakl et al
2002). Though this is not the most common method to form iron
nanoparticles, it has some advantages. This method can be scaled up with
ease. For the economical production of large quantities of nanoparticles,
milling is difficult to compete with. However, the product tends to be very
polydisperse in size and irregular in shape. Australia’s Advanced Powder
Technology Pvt. Ltd has successfully commercialized a wide range of
nanopowders by ball milling (Dallimore 1999).
2.11.2.2 Inert gas condensation
Inert gas condensation is the most versatile process in use today for
synthesizing experimental quantities of nanostructured powders. This
straightforward method is well suited for the production of metal
nanoparticles and has been widely accepted (Gleiter 1989; Sanchez-Lopez et
al 1997). Nanoscale iron particles were firstly synthesized through inert gas
condensation in 1989 (Gleiter 1989) and iron particles with an average
diameter of 17 nm were successfully prepared in 1997 by Sanchez-Lopez et al
(1997).
2.11.3 Chemical Synthesis Methods
2.11.3.1 Reverse micelle
Reverse micelle (microemulsion) synthesis begins by forming
inverse micelles containing the iron salt, then reducing the salt inside the
micelles, where the particle size is limited by the size of micelle (Li et al
67
2006b). This kind of synthesis offers an excellent method for preparing
nanoparticles with a very narrow size distribution and highly uniform
morphology (Carpenter 2001). This method has been commonly used for
noble metal synthesis but, because of its relative complexity, has not been
widely applied for the synthesis of iron nanoparticles.
2.11.3.2 Thermal decomposition
Research works on the synthesis of iron nanoparticles by thermal
decomposition were carried out in the late 1970s (Griffiths et al 1979; Smith
and Wychick 1980). The polymer-catalyzed decomposition of iron
pentacarbonyl (Fe(CO)5) to form iron nanoparticles in the range of 6 to 20 nm
was described by Smith and Wychick (1980). The thermal decomposition
method is frequently used in the formation of nanoparticles due to its ease of
use and lack of byproducts. The molecules are easy to decompose, but the
decomposition reactions are extraordinarily complicated. The decomposition
frequently changes rate and even order during the course of a reaction (Smith
and Wychick 1980). This is a complicating factor when attempting to grow
nanoparticles of controlled size and shape by this method.
2.11.3.3 Chemical vapor condensation
The chemical vapor condensation process was developed for
preparing a variety of materials, because a wide range of precursors is
commercially available (Wang et al 2003). The precursors can be a solid,
liquid or gas at ambient conditions, but are delivered to the reactor as a
vapour. Many applications of this method have been described over the last
20 years (Zaera 1989; Hahn 1997; Choi et al 2001).
68
2.11.3.4 Pulse electro-deposition
Recent research reveals that nanoscale iron particles can be
prepared successfully by pulse electro-deposition. When compared to
chemical reduction, sonochemistry and chemical vapour condensation,
electrodeposition is a cost effective fabrication technique with the ability to
“tailor” composition and properties of materials (Yoo et al 2007). The
synthesis method of nanoscale iron particles using a Fe anode and an inert Ti
cathode has been reported by Natter et al (2000).
2.11.3.5 Liquid chemical reduction
The basic idea of the liquid chemical reduction method is to add a
strong reducing agent into an ionic iron solution in order to reduce it to
nanoscale iron particles. Because of its simplicity and productivity, this
method is most widely used for the preparation of nanoscale iron particle.
This method was firstly used for NZVI synthesis in 1995 using sodium
borohydride as the reductant (Glavee et al 1995). This reductant was
successfully used for the formation of nanosized ZVI from both ferric
chloride and ferrous sulphate. Researchers have investigated the structure and
characteristics of NZVI formed by this method and obtained similar results
with regard to particle size and specific surface area (Choe et al 2000; Elliott
and Zhang 2001; Kanel et al 2005; Zhang 2003).
2.11.3.6 Other synthesis methods
Other than the synthesis methods discussed, there are some
alternative ZVI synthesis methods such as induction plasma methods and
microwave plasma method also used. However, they are not used
commercially because of the limitation in the availability of precursors and
the uneconomical nature of the process (Kalyanaraman et al 1998).
69
2.12 FENTON LIKE REACTIONS USING NANO ZERO VALENT
IRON (NZVI)
In heterogeneous Fenton reaction, oxidation of NZVI provides an
alternative means of inducing Fenton oxidation (Joo et al 2004) as shown in
Equation (2.28).
2220
2 2 OHFeHFeO (2.28)
Use of NZVI to induce Fenton oxidation has two advantages
beyond the addition of Fe2+/H2O2 (Bergendahl and Thies 2004):
NZVI is able to attach or coat on large particle, therefore NZVI
absorbed media could treat contaminated water passing through
a sand filter or other type of filtration system.
NZVI injected through wells could be immobilized in/on soil
grains in contaminated aquifers.
Tang and Chen (1996) investigated the effect of iron powder and
H2O2 on degradation of azo dyes and phenolic compounds. At pH below 2.5,
dye removal was found to be mainly by adsorption whereas, at higher pH, the
adsorption was less important. The addition of H2O2 doubled the dye removal
and was greater than 90%.
Takemura et al (1994) investigated the influence of steel wool, steel
foil, and reticulated iron in the presence of hydrogen peroxide on the
oxidation of perchloroethylene. The solids mentioned above were reduced the
perchloroethylene concentration from 100 mg/L to less than 0.1 mg/L in 24h.
The reaction with reticulated iron could be conducted at pH 5 – 9 with no
apparent iron-oxide by product.
70
Joo et al (2004) studied about the ability of NZVI to induce
oxidative degradation of contaminants. This study demonstrated that NZVI
was capable of degrading the herbicide molinate in the presence of oxygen. In
this study, it was shown that NZVI induced Fenton oxidation could degrade
over 60% of molinate at pH 8.1 and 65% at pH 4. Subsequent to this
investigation, Feitz et al (2005) showed that NZVI was capable of degrading a
wide range of organic compounds such as pesticides and herbicide that are
commonly found in agricultural run off. While NZVI appears to induce
oxidation effectively at low pH, the accumulation of ferric oxyhydroxides on
the surface of Fe0 in neutral to slightly alkaline environments is expected to
reduce the reactive surface area for the formation of hydroxyl radical and
hence, reduce the efficiency of NZVI remediation.
Bergendahl and Thies (2004) have investigated the feasibility of
oxidative removal of methyl-tert-butyl ether (MTBE) using NZVI. The results
showed that over 99% MTBE in water was degraded within 10 min and the
major oxidation product was acetone. The addition of H2O2 greatly enhanced
the degradation efficiency. The rate of degradation of MTBE increases with
an increase in H2O2: MTBE ratio.
Joo et al (2005) have observed the oxidation of organic compounds
in aqueous solution when mixed with nano scale zero valent iron particles in
the presence of air and no added H2O2. Evidently, H2O2 and Fe2+ are
generated by dioxygen corrosion of the metal and released into solution, or
react on the surface. This represents a novel use of zero-valent iron, normally
a reducing agent, for the oxidation reaction that has potential application for
soil treatment.
The iron minerals in combination with H2O2 were used to study the
natural attenuation of organic compounds in the environment, or to serve as
an alternative to Fenton type treatment (Tarr 2003). Although heterogeneous
71
oxidation processes is generally much slower than the corresponding solution
reaction at the same mole per liter of reagent concentrations, the
heterogeneous reactions are sometimes more efficient; that is, they consume
less peroxide per mole of contaminant degraded. Oxidation at the surface is
often speculated but there is little compelling evidence and further study is
clearly warranted.
2.13 FENTON LIKE REACTIONS USING IMMOBILIZED NZVI
NZVI was successfully used for the generation of hydroxyl radicals
in AOP system (Xu and Wang 2011; Zhang 2003; Zi et al 2006). Due to their
larger specific surface area and reactive sites, NZVI has gained prominence in
environmental remediation (Zhang 2003). Although NZVI particles were
successfully used in wastewater treatment, there are still some drawbacks
associated with the process and needs to be addressed. For example, NZVI
could coalesce into aggregates, which reduced the reactivity (Phenrat et al
2007) and filtration is required to remove NZVI particles at the end of the
treatment. To overcome these problems, NZVI particles were immobilized in
or on suitable solid supports, and also to expand the effective pH range of the
Fenton reaction. Some success in this regard was achieved with poly vinyl
alcohol (PVA) microsphere (Bai et al 2009). The others include calcium
alginate beads (Bezbaruah et al 2009), iron exchanged with Nafion
membranes (Maletzky et al 1999), iron modified clays (Feng et al 2003; De
Leon et al 2008) and with different iron modified supports such as silica
fabric (Bozzi et al 2002), Al2O3 (Muthuvel and swaminathan 2007), Zeolites
(Rios-Enriquez et al 2004), resins (Lv et al 2005) and cotton (Tryba 2008).
The main advantage of the heterogeneous catalyst is its separation from the
waste stream. The fundamental disadvantage is that the dissolved target
molecules must diffuse to the surface to reach active sites before they are
degraded.
72
Nafion is a perfluorinated oxyalkyl polymer with sulfonate groups
capable of binding cations. Ferric exchanged Nafion membrane is claimed to
be an effective photo-Fenton catalyst (Maletzky et al 1999; Dhananjeyan et al
2001). Fe-Nafion membranes appear to be effective up to an operational pH
of 5. Although it was possible to achieve mineralization of 2, 4
dichlorophenol at higher initial pH (up to 11), the pH always drifted quickly
to a value below 5 (Sabhi and Kiwi 2001). The detailed mechanism covering
the role of surface reaction and of adsorption/desorption of iron ions and
reactants in Fe-Nafion systems is unclear. Fernandez et al (1999) showed that
Nafion bound Fe(III) could be photo-reduced to Fe(II), which subsequently
undergoes the Fenton reaction. Some disadvantages of Nafion membrane for
practical application is their higher cost and these membranes are not totally
inert. Fluorinated radicals were found in solutions exposed to Nafion, which
are produced by photo-Fenton reaction (Bosnjakovic et al 2007).
The use of novel structured inorganic silica fabrics loaded with Fe
ions (Fe-EGF) by exchange–impregnation as a heterogeneous photo catalyst
was an alternative approach proposed to degrade organic compounds (Bozzi
et al 2002). It was proved that the degradation of oxalates on the Fe-Silica
fabric was due to iron ions leaching into solution that were re-adsorbed onto
the silica fabric when compounds were mineralized.
Ferric species were also immobilized on different types of ion
exchange resins and used as a heterogeneous catalyst for the degradation of
cationic and anionic dye pollutants under visible light ( >450 nm) irradiation
in the presence of H2O2. The cationic dyes are effectively photo degraded on a
cationic resin exchanged Fe(III) catalyst and the anionic dye is photo
degraded on an anionic resin supported catalyst (Lv et al 2005).
73
Ramirez et al (2010) studied the removal of orange II dye solution
using heterogeneous photo-electro-Fenton process using different iron
supporting materials. Amberlite and Purolite resins can be incorporated with
59 and 65 mg Fe/g of substrate, but a Nafion membrane can fix 45 mg/g of
supporting material. Iron desorption analysis indicates that more than 90% of
iron was retained by nafion membrane. Spectroscopic analysis and TOC
removal results indicate that the process can be carried out with any of the
iron supporting material used in this study.
The performance of a Fe3+-exchanged zeolite Y (Rios-Enriquez et
al 2004; Noorjahan et al 2005) and Fe-Zm5 zeolite (Kasiri et al 2008) as a
heterogeneous photo-Fenton catalyst was tested for the degradation of organic
compounds. Although some leaching of Fe3+ was observed, it amounted to
only few percent of the total iron present in the zeolites (0.3 - 2 mg/L in
solution) which did not play a major role in the reaction.
A common characteristic of designed heterogeneous catalyst for
photo- electro-Fenton reaction is that the optimal pH is generally 3. These
materials are, however, efficient at higher pH values and leaching is
minimized under these conditions. The contribution of heterogeneous
photo-electro-Fenton process depends on the kind of catalyst, experimental
conditions and target contaminants. In fact, the dissolved iron concentration is
time dependent and the change in optical properties of photo reactor in
absence of heterogeneous catalyst. An interesting phenomenon was observed
for most of proposed catalyst, that, the leached iron ions are found to re-
adsorb on catalyst surface and showed good catalyst stability.
An ideal supporting material for iron species is expected to have the
following attributes: (a) to favor strong surface chemical - physical binding
with the iron particles without affecting their reactivity; (b) to have high
specific surface area; (c) to have good adsorption capacity for the organic
74
compounds to be degraded; (d) to be in a physical configuration which favors
the ultimate liquid-solid phase separation; (e) to allow a reactor design that
facilitates the mass transfer processes; and (f) to be chemically inert. From the
immobilization point of view, a good adherence catalyst/support, and no
degradation of the catalyst activity by the attachment process are important.
The first condition is essential, since the support /catalyst junction should
resist to the stress derived from particle to particle and pesticide-fluid
mechanical interactions in the reactor environment, in order to avoid
detachment of catalyst particles from the support (Pozzo et al 1997).
PVA-alginate beads were used for the immobilization of NZVI in
this work. The PVA-alginate beads showed improved durability and strength
(Nunes et al 2010). Hydroxyl groups’ presence in PVA forms complexes with
metal ions (Kieber et al 2005). Hence, usage of PVA-alginate beads avoids or
alleviates the problems associated with the release of ferrous ions into waste
water (Bai et al 2009).