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).