Trends in the use of Fenton, electro-Fenton and photo-Fenton for the treatment of landfill leachate

Waste Management 30 (2010) 2113–2121

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Waste Management journal homepage: www.elsevier.com/locate/wasman

Review

Trends in the use of Fenton, electro-Fenton and photo-Fenton for the treatment of landfill leachate Muhammad Umar, Hamidi Abdul Aziz *, Mohd. Suffian Yusoff School of Civil Engineering, Engineering Campus, Universiti Sains Malaysia, 14300 Nibong Tebal, Penang, Malaysia

a r t i c l e

i n f o

Article history: Received 25 April 2010 Accepted 6 July 2010 Available online 1 August 2010

a b s t r a c t Advanced oxidation processes (AOPs) such as Fenton, electro-Fenton and photo-Fenton have been applied effectively to remove refractory organics from landfill leachate. The Fenton reaction is based on the addition of hydrogen peroxide to the wastewater or leachate in the presence of ferrous salt as a catalyst. The use of this technique has proved to be one of the best compromises for landfill leachate treatment because of its environmental and economical advantages. Fenton process has been used successfully to mineralize wide range of organic constituents present in landfill leachate particularly those recalcitrant to biological degradation. The present study reviews the use of Fenton and related processes in terms of their increased application to landfill leachate. The effects of various operating parameters and their optimum ranges for maximum COD and color removal are reviewed with the conclusion that the Fenton and related processes are effective and competitive with other technologies for degradation of both raw and pre-treated landfill leachate. Ó 2010 Elsevier Ltd. All rights reserved.

Contents 1. 2. 3.

4.

5. 6.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemistry of Fenton reagent . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Treatment by Fenton, electro-Fenton and photo-Fenton processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1. Fenton treatment. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2. Electro-Fenton treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3. Photo-Fenton treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of operating parameters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1. pH. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2. Fenton reagents dosage. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3. Reagent feeding mode. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.4. Temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.5. Initial COD . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.6. Effect of reaction time, current and distance between electrodes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.7. Recycling of Fenton sludge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Optimization. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1. Introduction Landfills serve as the ultimate solid waste disposal mechanism in countries worldwide (Umar et al., 2010). They fulfill the purpose * Corresponding author. Tel.: +60 4 5996215; fax: +60 4 5941009. E-mail address: [email protected] (H.A. Aziz). 0956-053X/$ - see front matter Ó 2010 Elsevier Ltd. All rights reserved. doi:10.1016/j.wasman.2010.07.003

2113 2114 2115 2115 2116 2117 2117 2117 2118 2118 2119 2119 2119 2119 2119 2120 2120

of controlled disposal of high quantities of solid waste at economical costs. The interaction of waste with water that percolates through the landfill produces highly polluted wastewater termed as landfill leachate. The properties of landfill leachate exhibit great temporal and site specific variations with concentration of contaminants ranging over several orders of magnitude (Deng and Englehardt, 2006). The composition of leachate depends on type

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of waste, amount of precipitation, site hydrology, waste compaction, cover design, interaction of leachate with environment, and landfill age, design and operation (Baig et al., 1999). The choice of treatment method is essentially based on the composition and properties of landfill leachate. For example, young leachates with high BOD5/COD ratio are effectively treated by biological methods. Biological treatment is commonly used for the removal of the bulk of leachate containing high BOD5 concentrations primarily due to its higher reliability, simplicity and cost effectiveness (Renou et al., 2008). Biological processes are driven by microorganisms which produce carbon dioxide (CO2) and sludge under aerobic and biogas under anaerobic conditions as the major by products. On the other hand, physicochemical processes are considered suitable as pre-treatment and/or full treatment for leachates with low BOD5/COD ratio and with high toxic constituents (Deng and Englehardt, 2006; Goi et al., 2010). Such processes include flocculation/precipitation (Aziz et al., 2007; Kurniawan et al., 2006; Tatsi et al., 2003), activated carbon adsorption (Foo and Hameed, 2009), chemical oxidation (Qureshi et al., 2002; Lopez et al., 2004), membrane filtration including reverse osmosis (RO) (Li et al., 2009; Ushikoshi et al., 2002), and nanofiltration (NF) (Alvarez-Vazquez et al., 2004; Robinson, 2007; Trebouet et al., 2001; Zhang et al., 2009). Physicochemical processes have been successfully applied for the removal of recalcitrant substances from stabilized leachate and refining the biologically pre-treated leachate. Among various physicochemical processes, advanced oxidation processes (AOPs) have been widely applied to enhance the biotreatability of wastewaters containing different organic compounds that are nonbiodegradable and/or toxic to microorganisms (Bila et al., 2005; De Morais and Zamora, 2005; Silva et al., 2004). A number of systems can be termed as AOPs and most of them use a combination of: two oxidants (O3 + H2O2), catalyst plus oxidant (Fe2+ + H2O2), oxidant plus irradiation (H2O2 + UV), oxidant plus photo-catalyst (H2O2 + TiO2 + hv), oxidants plus ultrasounds (US) (H2O2 + US) (Lopez et al., 2004). Typical AOP systems can be divided into homogenous systems with irradiation (O3/UV, H2O2/ UV, US, UV/US, electron beam, H2O2/US processes, H2O2/Fe2+/UV (photo-Fenton)), homogenous systems without irradiation (O3/ H2O2, O3/OH, H2O2/Fe2+ (Fenton’s)), heterogeneous systems with irradiation (TiO2/O2/UV, TiO2/H2O2/UV), and heterogeneous system without irradiation (electro-Fenton) (Huang et al., 1993). The homogeneous processes are known to possess lower mass transfer resistances between phases compared to heterogeneous ones, and therefore, favor the rapid degradation of pollutants (Wang et al., 1999). Recently, AOPs such as Fenton, electro-Fenton and photoFenton processes have been used to improve the quality of landfill leachate in terms of COD, color and odor removal (Deng and Englehardt, 2006). Fenton’s reagent has been used quite effectively for the treatment and pre-treatment of wastewater (Trujillo et al., 2006). Conventional Fenton process involves adding Fenton reagents (H2O2 and Fe2+) to the target wastewater. The formation of Fe3+ during the Fenton process results in the production of iron sludge because Fe3+ precipitates to iron oxyhydroxides particularly at higher pH. The resulting sludge is required to be treated and disposed of properly. Due to the high sludge produced during Fenton treatment, conventional Fenton process can be modified by the combined application of electricity i.e. electro-Fenton (Atmaca, 2009; Mohajeri et al., 2010; Zhang et al., 2006), and/or UV-light, i.e. photo-Fenton/photoelectron-Fenton (Altin, 2008; Kavitha and Palanivelu, 2004; Kim and Vogelpohl, 1998; Kim et al., 1997; Primo et al., 2008). In this study, the term ‘‘Fenton related processes” refers to electro-Fenton and photo-Fenton/photoelectro-Fenton processes. In addition to these, several other so called Fenton like processes (Cu2+/H2O2, UV/H2O2, UV/Cu2+/H2O2, UV/Fe2+/Cu2+/ H2O2, O3/H2O2, O3/OH) have been used for the degradation of organic compounds, but only conventional Fenton, electro-Fenton

and photo-Fenton have been reviewed in this study due to their relatively wider application for the treatment of landfill leachate. This review focuses on the state of the art in Fenton, electro-Fenton, and photo-Fenton treatment of landfill leachate and provides a comparative evaluation of these processes under various operating conditions as reported in literature by various authors. 2. Chemistry of Fenton reagent Hydroxyl radicals being one of the strongest oxidants (E = 2.73 V) are the main oxidizing species in the Fenton process. The Fenton reaction was first observed by Fenton (1984) and is based on an electron transfer between H2O2 and a metal ion such as ferrous iron (Fe2+) which acts as a catalyst. It is an economical method having no energy requirements as needed for the devices (ozonizers, UV lamps, and ultrasounds) in other AOPs (Lopez et al., 2004). Fenton and related reactions are potentially convenient ways to generate oxidizing species for pollutants degradation (Pignatello et al., 2006). Fenton process extends multiple benefits such as both iron and hydrogen peroxide are relatively cheap and safe, there is no mass transfer limitation except during coagulation where a high dosage of activator-ferrous salt is used and the process is technologically simple (Lopez et al., 2004; Pignatello et al., 2006). The mixture of H2O2 and Fe2+ produces hydroxyl radicals which are highly oxidative with respect to organic compounds present in the wastewater (Fenton, 1984). The fate of organic compounds and their degradation by products is primarily dependant on their reaction with hydroxyl radicals (Pignatello et al., 2006). Hydroxyl radicals attack the organic pollutants and lead to the complete destruction of contaminants to CO2, water and inorganic salts as end products. The classical Fenton process involves the sequence of following reactions (Deng and Englehardt, 2006; Pignatello et al., 2006).

Fe2þ þ H2 O2 ! Fe3þ þ  OH þ OH

ð1Þ

k1  70 M1 s1 ðRigg et al:; 1954Þ Fe3þ þ H2 O2 ! Fe2þ þ HO2 þ Hþ

ð2Þ

k2 ¼ 0:001  0:01 M1 s1 ðWalling and Goosen; 1973Þ 

OH þ H2 O2 ! HO2 þ H2 O

ð3Þ

k3 ¼ 3:3  107 M1 s1 ðBuxton and Greenstock; 1988Þ 

OH þ Fe2þ ! Fe3þ þ OH

ð4Þ

k4 ¼ 3:2  108 M1 s1 ðBuxton and Greenstock; 1988Þ Fe3þ þ HO2 ! Fe2þ þ O2 Hþ 3

k5 ¼< 2  10 M

1

s

1

ð5Þ ðDe Laat and Gallard; 1999Þ

Fe2þ þ HO2 þ Hþ ! Fe3þ þ H2 O2 6

k6 ¼ 1:2  10 M

1

s

1

s

1

ðBielski et al:; 1985Þ

2HO2 ! H2 O2 þ O2 5

k6 ¼ 8:3  10 M

ð6Þ ð7Þ

1

The generation of hydroxyl radicals (Eq. (1)) is very rapid. The net reaction (1)–(7) can overall be defined as the dissociation of H2O2 in the presence of iron as catalyst.

2Fe2þ þ H2 O2 þ 2Hþ ! 2Fe3þ þ 2H2 O

ð8Þ

The Eq. (8) implies that the reaction is completed under acidic conditions i.e. the presence of H+ ions is necessary for the decomposition of H2O2. Iron plays the role of catalyst in the above reactions by changing form between Fe2+ and Fe3+. However, in the Fenton chain reactions, the rate constant (k1) of Eq. (1) is 70 M1 s1, while that of Eq. (2) (k2) is 0.001–0.01 M1 s1, mean-

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ing that the rate of Fe2+ consumption is more rapid than rate of their generation. There are several reports on the reduction of Fe3+ to Fe2+ consuming H2O2 (Alegria et al., 2003; LipczynskaKochany, 1991; Pignatello et al., 2006), but the reduction of Fe3+ to Fe2+ is several orders of magnitude slower than the conversion of Fe2+ to Fe3+ in the presence of hydrogen peroxide (k2<<
RH þ HO ! H2 O þ R ! further oxidation k9 ¼ 107  1010 M1 s1

ð9Þ

Organic free radicals are formed as transient intermediates which are further oxidized to more stable products by ferric iron, oxygen, hydrogen peroxide, hydroxyl radicals (Hermosilla et al., 2009).

R þ H2 O2 ! ROH þ  OH

ð10Þ





R þ O2 ! ROO

ð11Þ

R þ Fe3þ ! Rþ þ Fe2þ 

2þ

R þ Fe



ð12Þ

3þ

! R þ Fe

ð13Þ

According to Bigda (1995), Fenton treatment is performed in the following four stages: pH adjustment, oxidation reaction, neutralization–coagulation, and precipitation, hence organic substances are removed by both oxidation and coagulation. Chemical coagulation in the Fenton process is associated with the formation of ferric hydroxo complexes.

½FeðH2 OÞ6 3þ þ H2 O $ ½FeðH2 OÞ5 OH2þ þ H3 Oþ ½FeðH2 OÞ5 OH

2þ

ð14Þ þ

þ H2 O $ ½FeðH2 OÞ4 ðOHÞ2  þ H3 O

ð15Þ

Within pH 3 and 7, the complexes changes to 2½FeðH2 OÞ5 OH2þ ! ½2FeðH2 OÞ8 ðOHÞ2 4þ þ 2H2 O 4þ

½2FeðH2 OÞ8 ðOHÞ2 

þ H2 O ! ½Fe2 ðH2 OÞ7 ðOHÞ3 

3þ

½2Fe2 ðH2 OÞ7 ðOHÞ3 

þ ½2FeðH2 OÞ5 OH

2þ

3þ

ð16Þ þ H3 Oþ

ð17Þ 5þ

$ ½3Fe2 ðH2 OÞ7 ðOHÞ4 

þ 5H2 O

ð18Þ The relative importance of coagulation and oxidation is a function of H2O2/Fe2+ ratio. According to Neyens and Baeyens (2003), chemical coagulation is dominant at lower H2O2/Fe2+ ratio, while chemical oxidation predominates at higher H2O2/Fe2+ ratios. The hydroxyl radicals produced at the start of Fenton reaction reacts mainly with Fe2+ because the reaction between Fe2+ and hydroxyl radicals is ten times quicker than the reaction between hydroxyl radicals and H2O2 (k4 = 3.2  108 M1 s1 and k3 = 3.3  107 M1 s1). Organic compounds present in leachate compete with the Fe2+ for hydroxyl radicals and thus affect the behavior of the Fe2+ (reaction (4) and (9)). 3. Treatment by Fenton, electro-Fenton and photo-Fenton processes 3.1. Fenton treatment Oxidation, neutralization, flocculation and sedimentation are the main steps involved in the Fenton process. Lowering the pH is an important step in the Fenton process because degradation of organic matter is most effective at pH  3 (Kochany and Lipczynska-Kochany, 2009). The process is generally carried out at ambient temperature. The samples to be analyzed are rapidly stirred at 80–400 rpm for 30 s to 60 min followed by increase in pH to neutral point (Deng and Englehardt, 2006). Neutralization is followed by flocculation prior to sedimentation. After sedimentation, COD of the supernatant is analyzed to measure the treatment performance. The measure of COD of the settled sludge gives the contribution of coagulation/flocculation towards removal of organics (Kang and Hwang, 2000). Leachate quality can be significantly improved in terms of organic content, color and odor by Fenton process (Deng and Englehardt, 2006). Table 1 shows the efficiency of Fenton process for the removal of COD as reported by various researchers. Table 2 shows the optimal pH and its control mode

Table 1 COD removal performance for different leachate types at various Fenton dosages. Leachate type

Initial COD (mg/L)

COD removal (%)

H2O2 (mg/L)

Fe2+ (mg/L)

Molar Ratio H2O2:Fe2+

Reference

Raw Raw Raw Raw Raw Raw Pre-treated biologically Pre-treated biologically Pre-treated by coagulation Pre-treated biologically Raw Raw Pre-treated by coagulation Pre-treated biologically Pre-treated by coagulation Anaerobically treated Mature Anaerobically treated Pre-treated biologically Pre-treated biologically – Pre-treated biologically Pre-treated biologically

1396, 2455 743 3000–4500 – 837, 1321, 6119 5700 ± 300 220 3300–4400 1100–1300 – 1000 10,540 22,400 – 1200–1500 1500 1800 1500 1500 338 2000 1100 2130

70–85% 60.9 60.8 75.1 <70 P 80 66 56 57 61 85 70 60 79 85 67.3 70 52 70 70 72 69 63 70

– 240 mM 20 ml/L – 0.075 M 650 5.4 mmol/L 5000 40 mM 5 g/L 0.075 M 10,000 2500 1200 1250 200 1500 200 1650 10 1500 900 200

– 4 mmol/L 20 g/L – 0.05 M 56 4.5 mmol/L 2000 40 mM 1.5 g/L 0.05 M 830 2500 1800 1000 300 2000 300 500 20 120 900 294

10.1 3 – 10.1 1.5 19.1 1.2 – 3 – 1.5 19.8 1.6 1.1 1.25 1.1 – – 0.8 – 1.6 1.1

Cotman and Gotvajn (2010) Cortez et al. (2010) Guo et al. (2010) Goi et al. (2010) Hermosilla et al. (2009) Kochany and Lipczynska-Kochany (2009) Wang et al. (2009) Primo et al. (2008) Deng (2007) Di Laconi et al. (2006) Zhang et al. (2005) Lopez et al. (2004) Pala and Erden, 2004 Gulsen and Turan (2004) Yoo et al. (2001) Lau et al. (2001) Kim et al. (2001) Wang et al. (2000) Kang and Hwang (2000) Welander and Henrysson (1998) Kim and Huh (1997) Bae et al. (1997) Gau and Chang (1996)

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used by various authors to treat landfill leachate. The efficiency of Fenton process varies between 52% (Kim et al., 2001) to 85% (Cotman and Gotvajn, 2010) for different types of leachates at various Fenton reagent dosages as shown in Table 1. The variance in COD

Table 2 pH value used in various studies and its control mode.

3.2. Electro-Fenton treatment

Leachate type

Optimal pH

pH control mode

Reference

Raw

3

Initial

Raw Raw –

3 3 3

Initial Initial –

Raw Raw

2.5 3.5

Initial Initial

Pre-treated biologically Pre-treated by coagulation – Raw Pre-treated biologically Pre-treated biologically Pre-treated biologically – Pre-treated by coagulation Anaerobically treated Mature – Mature

5 3 2.5 3.0 2.5 3–4 6.0 3–4.5 2.5–3 6.0 3.5 2.0–3.0 3.5

Initial Initial Constant Initial Constant Initial Initial – Constant Initial Constant Constant Initial

Cotman and Gotvajn (2010) Cortez et al. (2010) Guo et al. (2010) Ratanatamskul and Auesuntrachun (2009) Hermosilla et al. (2009) Kochany and LipczynskaKochany (2009) Wang et al. (2009) Deng (2007) Zhang et al. (2005) Lopez et al. (2004) Gulsen and Turan (2004) Lau et al. (2002) Lau et al. (2001) Kim et al. (2001) Yoo et al. (2001) Wang et al. (2000) Kang and Hwang (2000) Roddy and Choi (1999) Kim and Huh (1997)

Table 3 Biodegradability improvement after Fenton and related processes. Process

Initial BOD5/COD

Fenton Fenton Fenton

0.18 0.44 0.63

Photo-Fenton Fenton Fenton Electro-Fenton

0.13 0.2 – 0.1

Final BOD5/COD 0.38 0.68 0.88 0.4 0.5 0.5 0.3

removal can be attributed to several factors which are discussed in detail in the later section. Fenton process also improves the biodegradability and the Table 3 shows that the biodegradability can increase substantially after Fenton treatment. Fenton process has also been studied for the removal of color from landfill leachate and consistently higher color removals (>90%) have been reported in literature as shown in Table 4.

References Guo et al. (2010) Goi et al. (2010) Kochany and Lipczynska-Kochany (2009) De Morais and Zamora (2005) Lopez et al. (2004) Kim et al. (2001) Lin and Chang (2000)

Table 4 Color removal by Fenton and related processes.

In the electro-Fenton process, Fenton process and electro coagulation are combined to increase the degradability of organic compounds present in high strength wastewaters. A substantial increase in the oxidizing power of H2O2 occurs in the presence of electrically assisted Fenton process. The enhanced generation of hydroxyl radicals in the presence of electricity ensures considerable improvement in the removal of pollutants. Electro-Fenton process has two different configurations. In the first one, Fenton reagents are added to the reactor from outside and inert electrodes with high catalytic activity are used as anode material while in the second configuration, only hydrogen peroxide is added from outside and Fe2+ is provided from sacrificial cast iron anodes (Atmaca, 2009). Electro-Fenton process has been applied less frequently to treat landfill leachate than conventional Fenton process. Table 5 gives the COD removal efficiencies by electro-Fenton process as reported by various authors. There has been an increasing interest in using electro-Fenton process for landfill leachate and a couple of recent studied have been conducted by Mohajeri et al. (2010) and Atmaca (2009). Mohajeri et al. (2010) achieved higher COD (94%) and color removal (95.8%) than Atmaca (2009) who achieved 72% COD and 90% color removal. In the study by Mohajeri et al. (2010), the authors kept the distance between electrodes constant at 3 cm while Atmaca (2009) varied the distance between electrodes and found a range between 1.8 cm and 2.8 cm as suitable for maximum pollutant removal. Atmaca (2009) also took into consideration the removal of PO4–P and NH4–N and reported a removal of 87% and 26%, respectively under optimum conditions. The efficiency of electro-Fenton process can be further improved in the presence of UV irradiation by a process called photoelectro-Fenton. The catalytic effect of Fe2+ can be enhanced by assisting electroFenton process with UV irradiation. The photoelectro-Fenton process can increase the regeneration rate of Fe2+ in the presence of UV (Brillas et al., 1998). An increased concentration of OH increases the oxidative capability of the process (Peralta-Hernandez et al., 2006). In addition, H2O2 produces two OH by photocatalytic effect of UV irradiation (Chiou, 2007).

H2 O2 þ hm ! 2 OH

Process

Color removal (%)

Reference

Electro-Fenton Electro-Fenton Fenton Electro-Fenton Photo-Fenton/Fenton Electro-Fenton Fenton

95.8 90 >95 >90 >95 Complete 92

Mohajeri et al. (2010) Atmaca (2009) Wang et al. (2009) Altin (2008) Primo et al. (2008) Lin and Chang (2000) Kim and Huh (1997)

ð19Þ

The application of photoelectro-Fenton for the treatment of landfill leachate has been studied by Altin (2008). He compared the efficiency of photoelectro-Fenton process with electro-Fenton and reported an additional 10% COD removal by photoelectroFenton process in comparison to electro-Fenton. The author also reported higher than 90% color removal efficiency by electroFenton process. Lin and Chang (2000) studied electro-Fenton method together with chemical coagulation as pre-treatment

Table 5 Various operating parameters for electro-Fenton method and respective treatment efficiencies. H2O2/Fe2+ molar ration

H2O2 conc. (mg/L)

pH

Electrode distance (cm)

Reaction time (min)

COD removal (%)

Reference

1 – – 12 –

– 2000 2000 – 750

3 3 3 3 4

3 1.8–2.8 – 1.3–2.1 1.5

43 20 20 30–75 30

94.07 72 70 83.4 85

Mohajeri et al. (2010) Atmaca (2009) Altin (2008) Zhang et al. (2006) Lin and Chang (2000)

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M. Umar et al. / Waste Management 30 (2010) 2113–2121 Table 6 COD removal from landfill leachate by photo-Fenton process. Leachate type

Radiation (kW/m3)

Fe2+ mg/L

pH

Reaction time (min)

COD removal %

Reference

Pre-treated Raw Pre-treated Mature Pre-treated Pre-treated

– – – – 80 80

2000 10 – 56 70 56

3–3.5 2.8 3–4 3 2.6–3 3

60 60 30 – 40–70 120

86 57.5 >70 70 70–79 70

Primo et al. (2008) De Morais and Zamora (2005) Lau et al. (2002) Kim et al. (2001) Kim and Vogelpohl (1998) Kim et al. (1997)

biologically biologically biologically biologically

process before biological treatment. The authors used two pairs of anodic and cathodic electrodes (cast iron plates) while they added H2O2 to the electrolytic cell before the electrical current was initiated. In such arrangement certain amount of Fe2+ is dissolved into the leachate from the cast iron anode. The external addition of H2O2 results in the reaction of Fe2+ with hydrogen peroxide. Zhang et al. (2006) added the Fenton reagent from outside and used Ti/ RuO2 and IrO2 type electrodes as anode materials to treat high strength landfill leachate. The efficiency of electro-Fenton process can be reduced by the production of OH2 radicals which carries less oxidizing capacity than hydroxyl radicals according to Eqs. (3) and (4) (Mollah et al., 2001). Competitive electrode reactions are also considered important in reducing the efficiency of the electroFenton process (Zhang et al., 2006), but these interferences can be reduced by using proper Fe2+/H2O2 and Fe3+/H2O2 ratios and by controlling initial pH (Mollah et al., 2001). 3.3. Photo-Fenton treatment Photo-Fenton process has two main features: (a) the reduction of Fe3+ to Fe2+ to produce more hydroxyl radicals via photolysis (Deng and Englehardt, 2006; Kavitha and Palanivelu, 2004; Kim and Vogelpohl, 1998).

ðFe—OHÞ2þ þ hm ! Fe2þ þ OH

ð20Þ

and (b) the photo-decarboxylation of ferric carboxylates (Hermosilla et al., 2009; Kavitha and Palanivelu, 2004). i.e.

FeðIIIÞðRHCO2 Þ þ hm ! Fe2þ þ CO2 þ RH

ð21Þ

RH þ O2 ! RHO2 ! products

ð22Þ

As shown above, the reduction in amount of catalytic iron consequently reduces the final sludge volume, moreover, some additional organic compounds (carboxylates) can also be treated efficiently (Deng and Englehardt, 2006). Photo-Fenton process for landfill leachate has also not been studied as frequently as conventional Fenton process. Table 6 lists the studies related to photoFenton process and COD removal efficiencies under optimum conditions. Primo et al. (2008) reported photo-Fenton process as an efficient alternative for the treatment of biologically pre-treated landfill leachate and the authors attained 86% COD (Table 6) and total color removal under optimum conditions (Table 4). Kim and Vogelpohl (1998) and Kim et al. (1997) studied landfill leachate treatment with photo-Fenton process and the authors reported better treatment efficiencies. Kim et al. (1997) reported three times higher total organic removal (TOC) by assisting Fenton process with 80 kW/m3 UV radiation which enhanced the generation of hydroxyl radicals and consequently increased the degradation rate. According to Kim et al. (1997), the degradation rate increased six times at 80 kW/m3 radiation intensity while it doubled by increasing the radiation intensity from 80 to 160 kW/m3. Although several studies (Kim et al., 1997; Kim and Vogelpohl, 1998; Primo et al., 2008) have reported higher degradation of organics by photo-Fenton process, Hermosilla et al. (2009) realized no improvement in either COD or TOC removal of raw landfill leachate after coupling UV radiation with conventional Fenton process in both one step

dosing operation and by addition of H2O2 in batch mode every 20 min at a concentration of 0.075 M. The authors noted that, although, UV radiation helped in recycling the Fe2+ and consequently aided in the formation of additional hydroxyl radicals, the brown turbidity observed in the solution due to the higher amount of Fe2+ severely hindered the UV light transmission through the media and thus made UV ineffective in enhancing further degradation. It is worth mentioning that the effectiveness of UV irradiation can significantly vary for raw and pre-treated leachate due to the difference in the concentration of total dissolved solids and the level of turbidity. Additionally, the UV ineffectiveness can also be attributed to the variety of ferric complexes formed with each complex possessing different UV light absorbance capacity and consequently at the applied Fe2+ concentration (0.05 M), no significant improvement in either TOC or COD removal was recorded due to the formation of Fenton reaction intermediate products and decreased photoreduction of Fe3+ to Fe2+ (Hermosilla et al., 2009; Kim and Vogelpohl, 1998; Lopez et al., 2004). Carboxylic acids are the major intermediate by-products of Fenton reaction and are difficult to degrade by the Fenton process along with oxalates which are also known to be unreactive with hydroxyl radicals (Lopez et al., 2004). A decrease in UV light intensity decreases the photolysis of ferric-oxalate complexes and ultimately reduces the photo-regeneration rates, while a higher concentration of Fe2+ leads to quicker consumption of H2O2, hence decreases the production of hydroxyl radicals (Hermosilla et al., 2009). Although no increment in organics degradation was observed due to UV radiation, the amount of Fe2+ required was reduced 32 times by photo-Fenton process to obtain similar removal as for Fenton process which led to a decrease in the amount of final sludge from 25% to 1% (Hermosilla et al., 2009). The additional cost related to UV radiation in the photo-Fenton process can be compensated by reduced concentration of Fe2+ needed in photo-Fenton process together with decreased amount of sludge produced, which makes photo-Fenton process competitive with Fenton process in terms of overall treatment cost.

4. Effect of operating parameters 4.1. pH pH is one of the major factors that limits the performance of Fenton and Fenton related processes. It affects the speciation of iron and decomposition of hydrogen peroxide (Zhang et al., 2005). Efficiency of Fenton reaction is based on pH and acidic pH highly favors the oxidation reaction. The oxidation potential of hydroxyl radicals decreases with increase in pH from E0 = 2.8 V to E14 = 1.96 V (Kim and Vogelpohl, 1998). pH control in the Fenton process is carried out by using sulfuric acid and sodium hydroxide solution. The pH range to optimize the process is very narrow. According to Sedlak and Andren (1991), the production of hydroxyl radicals in the pH range 2–4 is higher because of either H2O2 regeneration or increase in reaction rates. Above pH 4, H2O2 decomposes in a different manner without any contribution to oxidation reactions (Zhang et al., 2005). Further, under alkaline

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conditions, H2O2 does not produce any hydroxyl radicals (Rivas et al., 2003b). pH control is either initial or continuous as shown in Table 2. pH values used in different studies are given in Table 2. In all the studies given in Table 2, pH range between 2 and 3.5 is used except by Lau et al. (2001), Wang et al. (2000) and Wang et al. (2009). Optimum pH value of 2.5 as determined by Zhang et al. (2005) is also an agreement with other studies on landfill leachate (Kang and Hwang, 2000; Lau et al., 2001), although some studies found maximum removal at pH slightly higher than 2.5 (Table 2). pH lower than optimum affects the pollutant removal by producing less hydroxyl radicals, increased scavenging effects of H+ and hydroxyl radicals (Tang and Huang, 1997), and termination of reaction between Fe3+ and H2O2. Higher pH affects the removal performance  by production of CO2 3 and HCO3 which scavenge hydroxyl radicals (Deng and Englehardt, 2006), decrease oxidation potential of hydroxyl radicals (Kim and Vogelpohl, 1998), deactivate Fe2+ by formation of ferric hydroxide (Bigda, 1995), reduce production of H+, and decomposition of H2O2 to water and oxygen (Kang and Hwang, 2000). More recently, Goi et al. (2010) studied direct Fenton treatment of raw landfill leachate without pH pre-adjustment. The authors found that the need for acidification during direct Fenton treatment can be avoided by increasing the dosage of H2O2. According to Goi et al. (2010), the pH decreased from initial 8.1 to 3.3 during 24 h Fenton treatment when a three step addition of H2O2 and a higher ratio of H2O2/COD (3/1) were used. The final pH was further decreased from initial 8.1 to 2.75 when both H2O2 and Fe2+ were added stepwise, however, prolonged Fenton treatment (24 h) is necessary to achieve lower final pH when preadjustment of pH is not implemented (Goi et al., 2010). pH increase during electro-Fenton process leads to electrocoagulation whereby pollutants are removed by electrostatic attraction and/or complexation of reactions due the conversion of Fe2+ and Fe3+ to Fe(OH)n type structures (Mollah et al., 2001). In the photo-Fenton process, the amount of photoregenerated Fe2+ is highly dependent on the pH. Although pH between 2.5 and 3.5 has been used in studies on photo-Fenton, Hermosilla et al. (2009) observed insignificant difference in COD removal for the pH range 2–4. During the photo-Fenton process, Kim and Vogelpohl (1998) observed the formation of scale on the immersion tube at pH 5 which reduced the transmission of radiation and hence, the photoreduction of Fe3+ complexes to Fe2+.

rate is dependent on the amount of catalyst (Fe2+) available, thus an adequate amount of Fe2+ is necessary for proper reaction initiation. An increase in Fe2+ dosage increases COD removal, although this increase is reported to be marginal above certain Fe2+ concentrations (Lin and Lo, 1997; Zhang et al., 2005, 2006). Kang and Hwang (2000) reported similar trends whereby they observed almost same COD removal beyond 500 mg/L Fe2+ concentration. According to Pérez et al. (2002), photochemical degradation process is known to be inhibited when excess Fe2+ ion is used because Fe2+ competes with the organics by the hydroxyl radicals as indicated in Eq. (4). Excess amount of Fe2+ produces extra amount of sludge and increase total dissolved solids and electrical conductivity (Gogate and Pandit, 2004). In the photo-Fenton process a higher Fe2+ can inhibit UV radiation penetration due to brown turbidity production and cause recombination of hydroxyl radicals (Kim et al., 1997). The amount of Fe2+ can be highly reduced in the photo-Fenton process because Fe2+ can regenerate by photolysis. In the absence of H2O2, the concentration of Fe2+ can increase to 30% by photolytic regeneration (Kim and Vogelpohl, 1998). An optimal ratio of H2O2 and Fe2+ is necessary to avoid scavenging effects and increased COD removal. The optimal ratio of H2O2 and Fe2+ fluctuates greatly as represented in Table 1 which is attributed to the type of pollutants present, matrix effect in complex wastewaters (Tang and Huang, 1997) and to the varying method of determining the optimal dosage (Deng and Englehardt, 2006). The ratio of H2O2/Fe2+ is required to be kept as low as possible to avoid recombination of hydroxyl radicals and reduce final sludge volume (Kim et al., 1997). The optimal doses are determined either (i) by varying Fe2+ concentration at a fixed random dosage of H2O2 and subsequently optimizing the H2O2 dosage at this Fe2+ concentration (Gulsen and Turan, 2004; Lau et al., 2001; Wang et al., 2000), (ii) by selecting the best combination of H2O2 and Fe2+ in terms of COD removal from various combinations (Pala and Erden, 2004), and (iii) by finding the optimal ratio of H2O2 and Fe2+ followed by the determination of optimal dosages at predetermined ratios (Lopez et al., 2004). Zhang et al. (2005) observed that the COD removal performance increased linearly with increase in H2O2/Fe2+ molar ratio by 1.5 of molar ratio. Increase in molar ratio beyond 2 produced less COD removals (Zhang et al., 2005). However, Deng (2007) and Cortez et al. (2010) established H2O2/Fe2+ molar ratio equivalent to 3 for mature leachate during Fenton oxidation as pre-treatment and post-treatment, respectively.

4.2. Fenton reagents dosage 4.3. Reagent feeding mode In the Fenton and Fenton related processes, the mass ratio of H2O2 and Fe2+ is very important in terms of overall cost and removal efficiency of the process. Excess or shortage of any of these two reagents results in the occurrence of scavenging reactions through Eqs. (3) and (4) (Lopez et al., 2004). Tang and Huang (1997) have shown that best oxidation efficiency is attained by reaction (9) when neither H2O2 nor Fe2+ is overdosed in order to make maximum hydroxyl radicals available for the oxidation of organic compounds. A decreased dosage of H2O2 does not generate enough hydroxyl radicals to achieve complete mineralization. Higher H2O2 dosages generally results in increase in percent degradation (Deng and Englehardt, 2006; Lin and Lo, 1997; Rivas et al., 2001), although the effectiveness of H2O2 is decreased when its dosage is increased beyond a certain point due to the formation of organic compounds which are difficult to be further oxidized (Kang and Hwang, 2000; Kochany and Lipczynska-Kochany, 2009). Excess H2O2 also results in iron sludge floatation or decreased sludge sedimentation because of the O2 off-gassing in response to autodecomposition of H2O2 (Kim et al., 2001; Lau et al., 2001). Table 1 lists the Fenton reagent dosages used in various studies and the respective COD removal efficiencies. Initial reaction

The addition of H2O2 at beginning or during the reaction imply changes in the ratios of H2O2/COD and H2O2/Fe2+ and ultimately the removal efficiency of COD (Primo et al., 2008). Fenton reagent addition mode has been studied by several authors (Deng and Englehardt, 2006; Hermosilla et al., 2009; Primo et al., 2008; Yoo et al., 2001; Zhang et al., 2005; Zhang et al., 2006). All these studies have recognized higher removal of COD (about 10%) by continuous addition of reagents. Zhang et al. (2006) termed this high removal to minimized effects of side reaction because of continuous addition of Fenton reagents. They argued that concentrations of Fenton reagents in the batch reactor are higher at initial stages because of their simultaneous addition, but their concentration reduces gradually as the reaction proceeds allowing undesired side reactions to occur which do not contribute towards oxidation. According to Turan-Ertas and Gurol (2002), keeping the concentration of H2O2 by stepwise addition reduces the hydroxyl radical scavenging, thus making more hydroxyl radical available for reaction with organic matter by decrease in the competency of reaction (3) with reaction (9) (Yoo et al., 2001; Zhang et al., 2005). The stepwise addition of Fenton reagents can also reduce the total chemical consumption.

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Yoo et al. (2001) observed a 25% reduction in Fenton reagents by stepwise addition while they also recorded additional 5% COD removal. An almost similar additional COD removal (4.5%) was achieved by Goi et al. (2010) when both H2O2 and Fe2+ were added in three steps in prolonged Fenton treatment (24 h). In treatment of landfill leachate by the photo-Fenton process, Primo et al. (2008) added Fe2+ at the beginning of reaction while added H2O2 in several doses during the process. The authors observed 8% increased COD removal when H2O2 was added at four doses. However, Wang et al. (2009) reported only a slight difference in COD removal for two and three step addition of Fenton reagents. Hence, they preferred two step addition of both H2O2 and Fe2+ with 1.5 h difference whereby an additional 11% COD was achieved. 4.4. Temperature Although temperature has a positive effect on the treatment efficiency in Fenton and related processes, the increase in organic removal due to temperature is relatively small compared to the other factors. Temperature lower than 8.3 °C results in slower initial kinetics (Deng and Englehardt, 2006), thus affects the reaction rate and the removal performance. On the other hand, temperature higher than 50 °C may negatively affect COD removal because the flocs may be destabilized at high temperatures. Because too low and too high temperature negatively impact the process efficiency, temperature between 20–30 °C can be considered as the most reasonable range because of relatively higher treatment efficiency in this temperature range. Hermosilla et al. (2009) observed a slight increase in COD removal when the temperature was increased from 25 to 45 °C in the Fenton process. Several other studies (Kang and Hwang, 2000; Rivas et al., 2003a; Zhang et al., 2005) have also reported similar results for the Fenton process. The authors observed increased COD removal with increase in temperature although percent COD removal decreased at temperatures higher than ambient. Although increase in COD removal because of temperature increase is smaller in comparison to the other factors, higher temperatures are beneficial for oxidation of organics (Zhang et al., 2005). 4.5. Initial COD COD is the amount of a specified oxidant that reacts with the sample under controlled conditions (APHA, AWWA, WEF, 2005). The determination of Fenton reagent’s dosages is to be made on the basis of initial COD for efficient treatment. The dosage of Fenton reagents can vary on the basis of nature of Fenton treatment i.e. pre-treatment or ultimate pre-treatment (Zhang et al., 2005). At higher initial COD, higher COD removal rates are reported for same amount of reagents used (Zhang et al., 2006). At constant H2O2/Fe2+ dosage, the authors reported that COD removal efficiency was 89.2%, 83.8%, 71.2% and 68.2% when the initial COD was 1000, 2000, 3000 and 4000 mg/L, respectively, whereas COD removal was 892, 1675, 2136, and 2726 mg/L, respectively (Zhang et al., 2006). This implies that, although the percent removal is higher when initial COD is low, the quantity of COD removal is higher at high initial COD values. A more recent study by Cortez et al. (2010) also reported higher COD removal at higher initial COD values. The authors reported that for initial COD of 743 and 93 mg/L, a removal of 231 and 57 mg/L was obtained, respectively, at H2O2 to Fe2+ molar ratio of 3 and reaction time of 40 min. 4.6. Effect of reaction time, current and distance between electrodes According to Lin and Chang (2000), the time to complete the oxidation reaction largely depends on the dosage of H2O2, hence the point of H2O2 consumption in the oxidation reaction can be

2119

termed as the point of oxidation termination. Zhang et al. (2006) studied effect of reaction time on COD removal efficiency by electro-Fenton process. The organic matter was degraded rapidly in the first 30 min of reaction time and later it gradually slowed down. Initial rapid degradation is largely due to the easily degradable organics. The current efficiency also increased in the first 15– 30 min but then it decreased gradually (Zhang et al., 2006). The decrease in current efficiency is because of the formation of hardly oxidizable products (Boye et al., 2003; Brillas et al., 2004). The higher DC current results in increased electro-regeneration of Fe2+ from Fe3+ but the treatment efficiency may not change significantly, therefore, the current density should be determined carefully to avoid additional energy consumption (Atmaca, 2009; Zhang et al., 2006). The treatment time required for the mineralization of pollutants can also be reduced by DC. Altin (2008) observed a drastic decrease in treatment time when the current was increased from 1.5 A to 2 A. According to Atmaca (2009), for the DC current higher than 2 A, speed of COD removal has been reported to decrease substantially while Zhang et al. (2006) also reported similar results although they used slightly higher current (2.5 A). In the electro-Fenton process, distance between electrodes is another important factor that affects the pollutants removal. According to Zhang et al. (2006), an optimum distance range could give additional COD removal. They reported that the COD removal efficiency remained same for electrodes distance between 1.3 and 2.1 cm and the shorter or larger distance achieved less COD removal. According to the Table 5, a range between 1.3 and 3 cm can be considered as the most suitable range for maximum COD removal. 4.7. Recycling of Fenton sludge The production of sludge is one of the main drawbacks of the Fenton process. However, Yoo et al. (2001) showed that the sludge can be recycled without adding further organic loading to the coagulation process. They found that the COD of the sludge prior to dewatering was only slightly higher than the effluent COD after coagulation process. Hence, the sludge can be recycled to reduce the consumption of coagulant and the final sludge volume. Additionally, a higher COD removal can be achieved by using sludge as alternative coagulant due to enhanced coagulation efficiency of recycled sludge (Yoo et al., 2001). Further, the recycling of sludge can increase the settling velocity of coagulated sludge due to the formation of relatively dense particle structures (Yoo et al., 2001). 5. Optimization Optimal reaction conditions in terms of cost and treatment efficiency are required to be established to improve the overall process performance. Two approaches are generally adopted for process optimization: changing one variable at a time (one-factor-at-a-time) to study the effects of variables on the response and two-level-factorial-design. The classical optimization technique of changing one variable at a time to study the effects of variables on the response has been widely used in process optimization (Zhang et al., 2009), but this classical optimization technique is time consuming and expensive, particularly for multivariable systems, and it does not shows the effect of interactions between different factors (Mohajeri et al., 2010). Two-level-factorial-design can be used to overcome the inter-variable interaction (Anderson and Whitcomb, 1996). Two-level-factorial-design offers certain advantages over the one-factor-at-a-time method. Twolevel-factorial design is a statistics based method that involves simultaneous adjustment of experimental factors at only two

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levels (Zhang et al., 2009). It can also consider the interactions among the variables and can be used to optimize the operating parameters. Although two-level-factorial design cannot investigate fully a wide range in the factor space, it can indicate major trends (Zhang et al., 2009). Numerous studies have been conducted to optimize the Fenton, electro-Fenton and photo-Fenton processes for the treatment of wastewater but studies on the optimization of landfill leachate are few. Zhang et al. (2009) evaluated the treatment performance of Fenton process in terms of COD removal with two-level-factorial design. The authors selected reaction pH, H2O2/ Fe2+ molar ratio, Fe2+ dosage and initial COD as variables in their two-level-factorial design. They found that the average effect on COD removal decreased as the pH, COD, and the interaction of pH and COD was increased, but an increase in Fe2+ dosage and H2O2/Fe2+ molar ratio increased the COD removal from landfill leachate of 3–5 years old landfill. Another recent study by Mohajeri et al. (2010) optimized the reaction conditions for landfill leachate by the electro-Fenton process. The authors determined optimum conditions in terms of pH, H2O2/Fe2+ molar ratio, current density and reaction time for COD and color removal from old landfill leachate by the electro-Fenton process. They reported 94% COD removal and 95.8% color removal under optimized conditions of pH 3, H2O2/Fe2+ molar ratio 1, current density 49 mA/cm2 and reaction time 43 min. 6. Conclusions The application of Fenton and Fenton related process to treat landfill leachate has received increased attention in the last decade. The efficiency of the Fenton process is highly reliant on reaction conditions and leachate composition. Appropriate molar ratio of Fenton reagents and initial pH are the two most important factors to achieve maximum COD removal performance. Overall, Fenton process is a promising technology for applications in landfill leachate treatment. It can achieve higher treatment efficiency than other physicochemical technologies noticeably coagulation and activated carbon adsorption. The process is also economical in comparison to other AOPs. Rivas et al. (2003a) estimated the operating costs roughly of 8  103 US$ per m3 of leachate and a ppm of COD removed. Contrary to membrane filtration and associated separation process, there is no mass transfer involved in this process thus pollutants are not transferred from one phase to another but are completely destroyed. Zhang et al. (2006) compared the cost of electro-Fenton and conventional Fenton process and according to the authors; electro-Fenton process cost Chinese Yuan Renminbi 18 (approximately equivalent to USD 2.64) for 1 kg COD removal while conventional Fenton process cost almost double than the electro-Fenton process, thus making electro-Fenton process more feasible both in terms of cost and treatment efficiency. There are certain drawbacks of the process namely production of sludge and operational and safety hazards associated with high acid requirements but these can be mitigated by choosing optimum quantities of Fenton reagents and implementing necessary safety measures. Considering the treatment cost of various processes, conventional Fenton process incurs high treatment cost due to greater requirement for chemicals i.e. Fe2+ and the disposal of final sludge, while photo-Fenton and electro-Fenton processes incurs higher equipment and energy requirements associated to the use of UV light and electricity. Hence, it is critical to define a set of conditions under which maximum removal of organics can be achieved by any of these processes and this can be accomplished by adjusting the amount of ferrous iron and hydrogen peroxide. Additionally, alternative processes such as photoelectroFenton shall be investigated further to achieve removal of phosphates in addition to COD and color.

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