Treatment of landfill leachate by the Fenton process

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Treatment of landfill leachate by the Fenton process Yang Deng, James D. Englehardt Department of Civil, Architectural and Environmental Engineering, University of Miami, McArthur Building Room 325, 1251 Memorial Dr., Coral Gables, FL 33146, USA

art i cle info

A B S T R A C T

Article history:

In recent years, studies of leachate treatment by conventional Fenton, photo-Fenton and

Received 21 October 2005

electro-Fenton processes have indicated that these methods can effectively reduce

Received in revised form

concentrations of organic contaminants and color. In addition, the process can increase

31 July 2006

the biodegradable fraction of organic constituents in leachate, particularly in mature or

Accepted 22 August 2006

biologically recalcitrant leachate. Oxidation and coagulation both play important roles in

Available online 11 October 2006

the removal of organics. Initial pH, dosages of Fenton reagents, aeration, final pH, reagent

Keywords:

addition mode, temperature, and UV irradiation may influence final treatment efficiency. In

Fenton treatment

this paper, current knowledge of performance and economics of Fenton processes for

Landfill leachate

treatment of landfill leachate as reported for laboratory, pilot and full-scale studies is

Oxidation

reviewed, with the conclusion that the Fenton process is an important and competitive

Coagulation

technology for the treatment or pretreatment of landfill leachate. & 2006 Elsevier Ltd. All rights reserved.

Contents 1. 2. 3. 4. 5. 6. 7.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3684 Design of Fenton reactors for leachate treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3685 Treatment efficiency . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3685 Fenton reaction versus Fenton-like reaction. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3686 Oxidation versus coagulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3686 Fenton-based treatment trains . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3686 Operating parameters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3688 7.1. Reaction pH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3688 7.2. Dosages of Fenton reagents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3690 7.3. DO. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3690 7.4. Effluent pH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3691 7.5. Addition modes of reagents and multiple-stage Fenton treatment. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3691 7.6. Recycling of iron sludge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3691 7.7. Temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3691 7.8. UV irradiation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3691

Corresponding author. Tel.: +1 305 284 2013; fax: +1 305 2843492.

E-mail address: [email protected] (Y. Deng). 0043-1354/$ - see front matter & 2006 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2006.08.009

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8.

1.

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Conclusion. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3691 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3692

Introduction

Sanitary landfills are the primary method currently used for municipal solid waste disposal in many countries, and leachate generated from landfills is a high-strength wastewater exhibiting acute and chronic toxicity. Untreated leachates can permeate ground water or mix with surface waters and contribute to the pollution of soil, ground water, and surface water. Additionally, leachate may cause malodors and aerosols, though these effects tend to be temporary and local (Christensen et al., 1992). The composition of municipal landfill leachate exhibits noticeably temporal and site-specific variation, with reported concentrations of contaminants ranging over several orders of magnitude. Biologically refractory organic constituents, ammonia, and heavy metals in leachate are three principal issues with regard to treatment and disposal. The organic content in leachate is usually described by chemical oxygen demand (COD), 5-day biochemical oxygen demand, or total organic carbon (TOC). Generally, high (3000–60,000, high BOD5/COD ratio (40.6), a high fraction of low-molecular organics characterize leachate from young landfills (o1–2 years old). In contrast, moderate COD (100–500(mg/L), low BOD5/COD ratio (o0.3), and a high fraction of high molecular-weight organics characterize mature leachate from old landfills (410 years old) (Calace et al., 2001; Tchobanoglous and Kreith, 2002). Ammonia, perhaps present in concentrations up to 2000 mg/L, may persist in the leachate with time, and may be toxic to biological processes for leachate treatment. Therefore, ammonia has been regarded as the most significant component in the leachate over the long term (Kjeldsen et al., 2002). Heavy metals are a concern due to their adverse effect on environment, though Kjeldsen et al. (2002) reported that most metals in leachate were at or below the US drinking water standards due to adsorption, precipitation and complexation in the landfill. Moreover, microbiological characteristics of leachate are less well known. Laboratory studies of biological and chemical treatment processes have been reported since the early 1970s (Boyle and Ham, 1974; Ho et al., 1974). Biological processes can be effective for young leachate with a high BOD5/COD ratio (Ehrig and Stegmann, 1992). However, these processes may not effectively treat leachate with a low BOD5/COD ratio, or with high concentrations of toxic constituents. Therefore, physicochemical processes are used for pretreatment or full treatment of such leachate. Processes include flocculation/ precipitation (Amokrane et al., 1997; Tatsi et al, 2003), activated carbon adsorption (Copa and Meidl, 1986), chemical oxidation (Ince, 1998; Qureshi et al., 2002; Lopez et al., 2004), membrane filtration including reverse osmosis (RO) (Ushikoshi et al., 2002) and nanofiltration (NF) (Trebouet et al., 2001). Recently, conventional, photo- and electro-Fenton processes have been investigated for landfill leachate treatment. In the Fenton process, hydrogen peroxide is added to wastewater in presence of ferrous salt, generating species that are

strongly oxidative with respect to organic compounds present (Fenton, 1894). Hydroxyl radicals (dOH) are traditionally regarded as the key oxidizing species in the Fenton processes (Barb et al., 1951; Yamazaki and Piette, 1991), though highvalence iron species and alkoxyl radicals (RO
) have also been proposed (Sheldon and Kochi, 1980; Rahhal and Richter, 1988; Bossmann et al., 1998; Buda et al., 2001). The classical Fenton free radical mechanism in the absence of organic compounds mainly involves the sequence of reactions below. Fe2þ þ H2 O2 ! Fe3þ þ d OH þ OH

(1)

Fe3þ þ H2 O2 ! Fe2þ þ HOd 2 þ Hþ

(2)

dOH þ H2 O2 ! HOd 2 þ H2 O

(3)

dOH þ Fe2þ ! Fe3þ þ OH

(4)

Fe3þ þ HOd 2 ! Fe2þ þ O2 Hþ

(5)

Fe2þ þ HOd 2 þ Hþ ! Fe3þ þ H2 O2

(6)

2HOd 2 ! H2 O2 þ O2

(7)

Hydroxyl radicals are rapidly generated through Eq. (1). In the above reactions, iron cycles between Fe2+ and Fe3+, and plays the role of catalyst. The net reaction of Eqs. (1–7) is the decomposition of H2O2 into water and O2 catalyzed by iron. 2H2 O2 ! 2H2 O þ O2

(8)

3+

2+

Although Fe can be reduced to Fe through Eq. (2), the rate is several orders of magnitude slower than that of Fe2+–Fe3+ conversion through Eq. (1). And the formed Fe3+ may precipitate to iron oxyhydroxides, particularly as pH is increased. Consequently, the undesirable iron sludge is generated, which needs proper treatment and disposal in application. In the presence of organic compounds, hydroxyl radicals can attack organics by four pathways: radical addition, hydrogen abstraction, electron transfer, and radical combination (SES, 1994). Carbon-centered radicals (Rd), formed in the reaction of hydroxyl radicals and organic compounds, may rapidly and usually irreversibly react with O2 in water. Rd þ O2 !! RðHþ Þ þ HOd 2

(9)

Rd þ O2 ! R2OOd !! R2Od

(10)

These radicals R , R–OO , and R–O may couple or disproportionate to form relatively stable molecules, or react with iron ions (Pignatello et al., 2006). The produced organic intermediates may continue to react with hydroxyl radicals and O2, thus leading to further decomposition and even final mineralization to water and CO2. Two energetically enhanced Fenton processes are applicable to treatment of high-strength organic wastewaters. First, the photo-Fenton process, employing ultraviolet (UV) radiation, can enhance reduction of dissolved Fe3+ complexes to Fe2+, and generate additional OH radical via photolysis. As a d

d

d

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result, the production of iron sludge can be reduced. However, the mechanisms of action including the roles of Fe2+ and Fe3+, as well as the equilibrium concentration of both species, are complex and not understood in detail (Gogate and Pandit, 2004). Second, the electro-Fenton process, in which either or both of H2O2 and Fe2+ can be generated electrochemically in situ, is an indirect electrochemical oxidation that employs OH radical generated by the Fenton reaction to oxidize organic compounds. Electro-Fenton designs used to detoxify hazardous organic compounds in water have been summarized by several authors (Qiang et al., 2003; Chang et al., 2004). In Fenton treatment of wastewaters, oxidation and coagulation by generated iron sludge both contribute to removal of organic constituents, though the effect of coagulation has not been well recognized. A detailed description of advanced oxidative reaction pathways occurring in Fenton treatment of wastewater was reported by Yoon et al. (2001). Hydroxyl radicals are responsible for oxidation. In spite of the indiscriminant and powerful nature of hydroxyl radical oxidation, Fenton oxidation is not efficient for oxidation of some organic compounds such as carbon tetrachloride, chloroform, tetrachloroethane, and trichloroethane that have been found from the leachates. (Bigda, 1995; Jimenez et al., 2002). On the other hand, chemical coagulation in Fenton systems is ascribed to the formation of ferric hydroxo complexes (Tang, 2003). The relative importance of oxidation and coagulation depends primarily on the H2O2/Fe2+ ratio. Chemical coagulation predominates at a lower H2O2/Fe2+ ratio, whereas chemical oxidation is dominant at higher H2O2/Fe2+ ratios (Neyens and Baeyens, 2003). Although the Fenton reaction has been applied to the destruction of toxic organics since the late 1960s (Huang et al., 1993), the approach has been reported for leachate treatment recently. Current state-of-the-art in Fenton treatment of landfill leachate, reported for laboratory, pilot and full-scale studies, is reviewed in this paper.

2. Design of Fenton reactors for leachate treatment Lab and pilot scale experimental procedures for Fenton treatment of leachate generally include steps for oxidation, neutralization, flocculation and solid–liquid separation. Lab scale experiments have been operated in the batch mode, except by Roddy and Choi (1999) who investigated continuous flow mode. Because low pH favors Fenton oxidation, the pH must be lowered. Generally, two pH control methods are used. First, only the initial pH is adjusted. Second, pH is adjusted initially and controlled continuously at the desired value with a pH controller using sulfuric acid and sodium hydroxide solution. Fenton oxidation is initiated by addition of Fenton reagents, typically ferrous sulfate and H2O2. Rapid mixing continues for from 30 s to 60 min at impellor speeds of 80–400 rpm. Subsequently, final pH is increased to the neutral range with sodium hydroxide or lime solution. During this procedure large amounts of iron sludge may form, because Fe3+ is converted to ferric-hydroxo complexes. Subsequently, paddle flocculation for 10–30 min at 20–80 rpm is performed, followed by sedimentation for periods of from 30 min to

3685

several days. COD, BOD5, and TOC in the supernatant are measured to evaluate overall treatment efficiency, and COD in the settled sludge can be used to evaluate the contribution of coagulation/flocculation to the removal of organics (Kang and Hwang, 2000). Several operational problems may lead to lower COD removal efficiency in the field than is found in laboratory tests (Roddy and Choi, 1999; Zhang and Huang, 2001). The primary reported problem is foaming due both to CO2 produced from carbonate species at acidic pH and to organic foaming agents in the leachate. Foaming takes place in the mixing and oxidation tanks, leading to the requirement for a much larger tank volume. Other problems hindering operations include malfunctioning pumps, controllers and flowmeters. Bigda (1995) also reported that much less iron salt was generally needed in full-scale tests than in lab tests, for treatment of industrial wastewater. Information on commercial full-scale Fenton reactors for landfill leachate treatment is limited in the literature. However, commercial Fenton reactors for industrial wastewater treatment have been used for the past 10–15 years, and they have been reviewed by several authors (Bigda, 1995, 1996; Tang, 2003; Gogate and Pandit, 2004). The reactors are typically operated in a batch mode with great flexibility, and the process includes oxidation, neutralization, flocculation and solid–liquid separation. Air sparging may be used to remove volatile organic compounds (VOCs) in solution. Such reactors are non-pressurized and stirred, with metering pump for the addition of acid, base, and Fenton or Fentonlike reagents. The pH sensor/controllers are used with acid and base addition to maintain a constant desired pH in reactors, and redox potential sensor/controllers are used to control the addition of Fenton reagents. Level sensors control the feed pump, to prevent overflow and accidental discharge (e.g., if the operator leaves the bottom valve open when filling). The interior of the reactor vessels must be acidresistant. During oxidation, the order of addition of wastewater and chemicals is important. First, the vessel is filled with wastewater. Second, pH is adjusted with dilute acid to prevent formation of iron oxyhydroxides. Third, iron salt is added. The pH depression caused by iron salt addition is controlled by a pH sensor/controller. Fourth, H2O2 is introduced slowly. During neutralization, lime is used to raise the pH to 6–9, form iron sludge, decrease total dissolved solids (TDS), and precipitate metals. Subsequently, coagulation agents are metered to support flocculation, and solids are separated by sedimentation and filtration.

3.

Treatment efficiency

Leachate quality in terms of organic content, odor, and color can be greatly improved following Fenton treatment. Most important, the Fenton process can significantly remove recalcitrant and toxic organic compounds, and increase the biodegradability of organic compounds. Reported COD removal efficiencies range from 45% (Kim et al., 2001) to 85% (Roddy and Choi, 1999), and reported final BOD5/COD ratio can be increased from less than 0.10 initially to values ranging

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from 0.14 to more than 0.60 (Kim et al., 1997; Lau et al., 2002), depending on leachate characteristic and dosages of Fenton reagents. Color and odor in leachate can also be reduced considerably. Kim and Huh (1997) reported decolorization efficiency as high as 92% in Fenton treatment of a mature leachate. And, Lin and Chang (2000) found that leachate after electro-Fenton treatment was colorless and odorless, though almost 300 mg/L of the initial COD remained. However, little ammonia is reduced, and even TDS may be increased subsequent to Fenton reaction. In spite of having significantly strong oxidizing capacity, hydroxyl radicals produced in Fenton oxidation cannot oxidize ammonia. Consequently, almost all of ammonia in leacahtes can go through Fenton oxidation reactors (Lin and Chang, 2000; Lau et al., 2002; Deng, 2006; Englehardt et al., 2006). Englehardt et al. (2006) reported that TDS in effluent increased by 100% compared with that in influent in treating a mature leachate with Fenton method. The increase is due to input of the acid used to adjust pH, and the iron salts. And information on the removal of other constituents (e.g., metal ions) from leachate by Fenton treatment is scarce.

4.

Fenton reaction versus Fenton-like reaction

Fenton reaction initiated by Fe2+ and H2O2, and Fenton-like reaction initiated by Fe3+ and H2O2, have been used for treatment of leachates. From a mechanistic standpoint, it is meaningless to differentiate the both processes, because Fe2+and Fe3+ are in the chain of Fenton reactions (Eqs. (1–7)). Once Fenton oxidation is initiated, all initially added Fe2+ are rapidly oxidized to Fe3+, such that the system behaves regardless of the initial oxidation state of iron (Pignatello et al., 2006). However, a great difference in practice is that a greatly quick generation of hydroxyl radicals may occur at the beginning of Fenton oxidation, whereas Fenton-like oxidation has a slow generation rate of hydroxyl radicals. The reason is that the rate constant in Eq. (1) (the main reaction at the initial stage of Fenton oxidation) is much higher than that in Eq. (2) (the main reaction at the initial stage of Fenton-like oxidation), such that the latter reaction becomes a ratelimiting step and slows down the release of hydroxyl radicals. Few investigations have been carried out to compare the both processes for leachate treatment, and the reported results are controversial. Rivas et al. (2003a) reported that Fenton and Fenton-like reactions had similar organics removal efficiencies. However, Kim et al. (2001) found that Fenton reaction achieved a higher COD removal and a higher BOD5/COD than Fenton-like reaction. And the optimal pH 3.0 for Fenton oxidation was below the optimal pH 4.5 for Fenton-like reaction.

5.

Oxidation versus coagulation

Few studies have been conducted to characterize the relative removal of organic constituents by oxidation as opposed to coagulation, through Fenton treatment of landfill leachate. Yoon et al. (1998) found that organic removal rates by Fenton treatment and by simple coagulation were similar, in that

high molecular organics were removed more readily and selectively. Therefore, they concluded that coagulation played a primary role in the removal of organics from leachate by Fenton treatment. Wang et al. (2000) and Lau et al. (2001) reported that oxidation and coagulation were responsible for approximate 20% and 80% of overall COD removal, respectively, in Fenton treatment of a biologically stabilized leachate. Kang and Hwang (2000) found that pH, and Fenton reagent dosages, significantly affected removal of COD from leachate by oxidation and coagulation. At a maximum, removal of COD by oxidation was estimated at 2–3 times removal by coagulation. And, overall COD removal efficiency and COD removal efficiency by oxidation linearly increased with increasing H2O2 concentration. However, this increase in COD removal efficiency by oxidation would reach saturation at higher H2O2 concentrations due to the presence of recalcitrant organics (Deng, 2006; Englehardt et al., 2006).

6.

Fenton-based treatment trains

In practice, the selection of an appropriate scheme for Fenton treatment of leachate is a complicate process that involves consideration of many interrelated factors, including required treatment efficiency, ultimate disposal, capital and operating costs, and operational complexity. Several flow sheets reported for Fenton-based leachate treatment schemes are shown in Fig. 1. These flow sheets can be roughly grouped into four categories: direct Fenton treatment of raw leachate, Fenton pretreatment prior to biological treatment, Fenton treatment preceded by physical/chemical treatment, occasionally followed by biological treatment, and, Fenton treatment preceded by biological treatment, occasionally followed by physical/chemical or biological treatment. The difference between direct Fenton treatment (Fig. 1(a)), and Fenton pretreatment for a biological process (Fig. 1(b)), is that the former focuses on COD removal, while the latter focuses on increasing the BOD5/COD ratio. The first design scheme (Fig. 1(a)) is appropriate for mature leachate, because biodegradable organics in young leachate can be economically removed by biological processes. The second scheme (Fig. 1(b)) is also appropriate for mature leachate. Lopez et al. (2004) found that dosages of Fenton reagents required to achieve an exploitable improvement in organic biodegradability are significantly lower than those required to achieve a minimal COD in a raw leachate. The same authors utilized conventional Fenton to increase the BOD5/COD ratio of from 0.22–0.50. Kim and Huh (1997) also raised the BOD20/COD ratio in a mature leachate from 0.10 to 0.58 by a conventional Fenton process. Kim et al. (1997) used photo-Fenton treatment to increase the BOD5/COD ratio of leachate from o0.05 to 40.60. However, it should be emphasized that final BOD5/COD ratio cannot be increased to 40.50 in some cases. This result may be attributed either to the refractory organic fraction in the leachate, or to inadequate dosages of Fenton reagents. Under such circumstances, higher dosages of Fenton reagents, another scheme for Fenton treatment, or another treatment technology should be considered.

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4 0 (200 6) 368 3 – 369 4

Fenton (a) Fenton

Biological

(b) Physical/Chemical

Fenton

Physical/Chemical

Fenton

Coagulation (Fe 3+)

Fenton

(c)

Biological

(d)

Iron sludge (e) Biological

Fenton

Biological

Fenton

Biological

Fenton

(f)

Physical/Chemical

(g)

Biological

(h) Fig. 1 – Flow sheets for Fenton treatment of landfill leachate.

The third process design (Figs. 1(c) and (e)) frequently involves coagulation as pretreatment, to reduce organic loading on the Fenton process. Since the 1970s, lime, alum and ferric salts have been extensively investigated for full treatment and pretreatment for landfill leachate (Thornton and Blanc, 1973; Cook and Foree, 1974; Ho et al., 1974; Millot et al., 1987; Christensen et al., 1992; Amokrane et al., 1997; Forgie, 1988; Kim et al., 2001; Trebouet et al., 2001; Wang et al., 2002; Tatsi et al., 2003; Monje-Ramirez and Velasquez, 2004; Pala and Erden, 2004; Silva et al., 2004; Wu et al., 2004). COD removal efficiencies reported for coagulation depend primarily on coagulant species, coagulant dose, pH and leachate characteristics, ranging widely from 10% to 80%. Several workers reported that coagulation favored removal of high molecular organic compounds in leachate (Chian and DeWalle, 1976; Slater et al., 1985; Yoon et al., 1998). A modification of coagulation (ferric salt) combined with Fenton treatment involves the recycle of Fenton sludge, as shown in Fig. 1(e). In this scheme, a portion of the Fenton sludge is returned to the coagulation tank to increase COD removal efficiency, and reduce coagulant consumption and sludge disposal cost (Yoo et al., 2001). Finally, if effluent from the Fenton unit contains COD above the discharge standard, posttreatment may be employed (Lin and Chang, 2000).

According to the fourth treatment scheme (Fig. 1(f–h)), biological pretreatment including one or more aerobic/ anaerobic processes is used to reduce biodegradable organics or ammonia at a low operating cost (Gau and Chang, 1996; Bae et al., 1997; Yoon et al., 1998; Wang et al., 2000; Lau et al., 2001; Gulsen and Turan, 2004). Gau and Chang (1996) used activated carbon adsorption as a post-treatment to decrease COD from 320 to 150 mg/L. The reason for this efficiency may be that the effluent from Fenton leachate treatment contains a high fraction of low molecular organics (Yoon et al., 1998), which may be efficiently removed by activated carbon adsorption (Chiang et al., 2001). Another option for reduction of COD in Fenton-treated leachate is activated sludge treatment (Bae et al., 1997). In addition to the process designs described above, twostage Fenton treatment processes, and biological treatments of mixtures of Fenton effluent and municipal wastewater, have been reported. The two-stage Fenton process, comprised of two successive Fenton treatment units, can be used in place of a single Fenton treatment unit. Gau and Chang (1996) reported that the second Fenton step could further reduce COD in leachate, though dosages of Fenton reagents in the second stage were much higher than those in the first step.

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Control mode of pH: constant, pH is continually controlled during the whole oxidation stage; initial, only initial pH is adjusted to the desirable value. N/A: not available b

a

Lab scale Pilot scale Lab scale Lab scale Pilot scale Pilot scale Lab scale N/A Lab scale Lab scale Lab scale Lab scale Lab scale N/A Lab scale Lab scale Batchwise Batchwise Batchwise Batchwise Batchwise Batchwise Continuous flow N/A Batchwise Batchwise Batchwise Batchwise Batchwise N/A Batchwise Batchwise

Conventional Fenton Photo-Fenton like Conventional Fenton Conventional Fenton Photo-Fenton Photo-Fenton Conventional Fenton Conventional Fenton Conventional Fenton Conventional Fenton Photo-Fenton Electro-Fenton Conventional Fenton Conventional Fenton-like Conventional Fenton Conventional Fenton

Experiment scale

N/Ab Pretreated biologically Pretreated biologically N/A Pretreated biologically Pretreated biologically N/A Mature Mature Mature Pretreated biologically Mature Pretreated biologically Mature Pretreated biologically Pretreated biologically Constant Initial Constant Constant Initial Initial Constant N/A Constant Initial Initial Initial N/A N/A Initial Initial 2.0–3.0 2.5 2.5 2.5 3.0 3.0 3.0 3.0 3.5 3.5 3.0–4.0 4.0 4.0 4.5 6.0 6.0

An essential characteristic of the Fenton process is that pH in the acidic range strongly favors oxidation. Optimal pH values reported for conventional, photo-and electro-Fenton processes for landfill leachate treatment range between 2.0 and 4.5, except those of Wang et al. (2000) and Lau et al. (2001) (pH 6.0), as shown in Table 1. A pH below optimal can inhibit oxidation in three ways. First, at extremely low pH values, the [Fe(H2O)]2+ formed reacts relatively slowly with H2O2, producing less OH radical (Gallard et al., 1998). Second, the scavenging effect of H+ on dOH becomes more important at a lower pH (Tang and Huang, 1996). Third, exceptionally low pH can inhibit reaction between Fe3+ and H2O2 (Pignatello, 1992). On the other hand, a pH above optimal, especially in the neutral-to-alkaline range, also hinders Fenton oxidation. Five mechanisms for this inhibition have been suggested. First, the absence of H+ can inhibit the decomposition of H2O2 to reduce production of dOH (Walling, 1975). Second, H2O2 itself rapidly decomposes to water and oxygen with increasing pH above 5 (Meeker, 1965). Third, the ferrous catalyst is deactivated with the formation of ferric oxyhydroxide at pH above 5 (Bigda, 1995). Fourth, under neutral and alkaline conditions, the primary forms of aqueous carbonate system and HCO are CO2 3 3 , both well-known dOH scavengers. Finally, the oxidation potential of OH radical decreases with increasing pH from E0 ¼ 2.8 V at pH 0 to E14 ¼ 1.95 V at pH 14 (Kim and Vogelpohl, 1998). In the photo-Fenton process, two reasons not previously mentioned strongly favor oxidation process at pH 2–4.5. First, at around pH 3.0, highly soluble Fe(OH)2+ is the predominant ferric hydroxide complex as opposed to free Fe3+, Fe(OH)+2 , and Fe2(OH)4+ 2 , which are less photoreactive (Faust and Hoigne, 1990). Second, amorphous iron oxyhydroxide sludge that may accumulate at pH above 5.0 prevents the transmission of UV light through the reactor (Kim et al., 1997; Kim and Vogelpohl, 1998). In Table 1, a relatively high pH (6.0) was reported to be the optimal initial value in two cases. The reason may be attributed to great suppression of organics removal by coagulation. In the both cases, coagulation played a much important role than oxidation, and the ratio of COD removal

Fenton type

Reaction pH

Operation mode

7.1.

Leachate characteristic

Operating parameters

Table 1 – Optimal pH in conventional, photo-Fenton, and electro-Fenton treatment of landfill leachate

7.

Reference

As mentioned earlier, although Fenton treatment can increase the biodegradability of organic compounds, the increase may occasionally be inadequate to support subsequent biological treatment due to the presence of organics that are refractory to Fenton oxidation. For example, Kim et al. (2001) raised leachate BOD5/COD from 0.14 to 0.22 and 0.27 by Fenton and Fenton-like methods, respectively. Lau et al. (2002) applied photo-Fenton treatment to increase leachate BOD5/COD from 0.08 to 0.14. And, Lin and Chang (2000) obtained a BOD5/COD ratio of 0.30 by electro-Fenton treatment of a leachate (original BOD5/COD ¼ 0.10). To remedy this deficiency, the leachate effluent can be mixed with municipal wastewater prior to biological treatment (Lin and Chang, 2000).

Roddy and Choi (1999) Kim and Vogelpohl (1998) Gulsen and Turan (2004) Zhang et al. (2005) Kim and Vogelpohl (1998) Kim et al. (1997) Roddy and Choi (1999) Kim et al. (2001) Kang and Hwang (2000) Kim and Huh (1997) Lau et al. (2002) Lin and Chang (2000) Gau and Chang (1996) Kim et al. (2001) Wang et al. (2000) Lau et al. (2001)

40 (2006) 3683– 3694

pH control modea

WAT E R R E S E A R C H

Optimal pH

3688

22,400 2130

Conventional Fenton

Conventional Fenton

Pretreated by coagulation Pretreated biologically

a

N/A: not available.

Electro-Fenton

Mature

Photo-Fenton-like

Pretreated biologically

Photo-Fenton

Photo-Fenton

Pretreated biologically

Fenton-like

Pretreated by coagulation Pretreated biologically

Photo-Fenton

1150

Conventional Fenton

Pretreated biologically

950

513

1150

440

1100

338

750

800

2438

660

1150

34,000

900

10

N/A

N/A

56

31

72

558

900

20

1800

1500

1500

184

294

2500

830

300

Fe2+ (Fenton), Fe3+(Fentonlike) (mg/L)

N/A

N/A

1

35.0

26.2

100

1.6

0.8

1.1

1.6

0.7

8.9

1.1

1.6

19.8

1.1

Molar H2O2:Fe2+

68

71

70

80

70

80

63

72

85

45

69

50

70

79

60

70

COD removal efficiency

1.15

2.19

3.03

1.88

1.43

12.0

1.3

0.04

N/A

1.85

0.64

3.13

0.13

0.14

1.58

0.19

H2O2 (mg/mg COD)

N/A

N/A

0.07

0.09

0.09

0.2

1.3

0.08

N/A

1.85

1.60

0.58

0.20

0.14

0.13

0.29

Fe2+ (Fenton) Fe3+(Fentonlike) (mg/mg COD)

Consumed reagent

Lau et al. (2001) Lopez et al. (2004) Pala and Erden (2004) Gau and Chang, (1996) Gau and Chang (1996) Kim et al. (2001) Kim et al. (2001) Gulsen and Turan (2004) Welander and Henrysson (1998) Bae et al. (1997) Rivas et al. (2004) Kim and Vogelpohl (1998) Kim and Vogelpohl (1998) Kim et al. (2001) Lau et al. (2002) Lin and Chang (2000)

Reference

WAT E R R E S E A R C H

Mature

3530

Conventional Fenton

Pretreated biologically

1200

Conventional Fenton

Pretreated biologically

N/Aa

Fenton-like

Pretreated biologically

600 1500

1800

Conventional Fenton

1000

200

2500

10,000

200

H2O2 (mg/L)

Reagent dosages

1800

640

Conventional Fenton

After one-stage Fenton Pretreated biologically

10,540

Conventional Fenton

Raw

1500

COD (mg/L)

Initial

Conventional Fenton

Fenton type

pretreated biologically

Leachate characteristic

Table 2 – Comparison of Fenton reagent dosages and COD removal in Fenton treatment of landfill leachate

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by coagulation to by oxidation could reach 4:1. At pH 4.0–6.0, the colloidal organics mostly carried negative charges due to dissociation of acidic functional groups. However, at pHo4.0, these negative charges were neutralized because these functional groups became undissociated. As a result, the coagulation was severely inhibited, such that overall organics removal was decreased to some extent at pHo4.0.

7.2.

Dosages of Fenton reagents

Dosages of Fenton reagents determine, to a large extent, the operating costs and efficiency of organics removal for the process. Determining optimal dosages involves consideration of both absolute levels and their relative ratio. Generally, removal of organics increases with increasing concentration of iron salt. However, the removal increment may be marginal when the concentration of iron salt is high. A similar trend is observed for H2O2. Also, excess iron salt contributes to an increase in effluent TDS and electrical conductivity, as well as in the amount of iron sludge that requires treatment (Gogate and Pandit, 2004). Excess H2O2 results in iron sludge flotation, due to O2 off-gassing caused by autodecompostion of excess H2O2 (Kim et al., 2001; Lau et al., 2001), and residual H2O2 may inhibit downstream biological treatment. Dosages of Fenton reagents and COD removal efficiencies reported for Fenton treatment of leachate are shown in Table 2. The large discrepancies in reported optimal ratios of H2O2 to Fe2+, and in reported COD removed per unit mass of Fenton reagents, are ascribed to variations in leachate quality and to differing methods of determining the optimal dosage. Three methods have been used to optimize addition rates. First, Fe2+ dosage has been varied at a fixed arbitrary dosage of H2O2, followed by optimization of the H2O2 concentration at this Fe2+ dosage (Wang et al., 2000; Lau et al., 2001; Gulsen and Turan, 2004). Second, the best among several combinations of Fenton reagents’ dosages has been selected as optimal in terms of COD removal over the tested range of H2O2 and Fe2+ concentrations (Pala and Erden, 2004). Third, the optimal ratio of H2O2 to Fe2+ is first found, and optimal dosages of Fenton reagents are then determined at this ratio (Lopez et al., 2004; Zhang et al., 2005). The success of all three methods is dependent on the extent of iteration. Understanding the roles of H2O2 and iron in the removal of organic compounds by Fenton treatment helps in determining the optimal reagent dosages. Of the two reagents, H2O2 is more critical because it directly affects the theoretic maximum mass of dOH generated. Therefore, H2O2 dosage depends heavily on initial COD with high initial COD generally requiring more H2O2. The theoretical mass ratio of removable COD to H2O2 is 470.6/1000. That is, 1000 mg/L H2O2 theoretically removes 470.6 mg/L COD by oxidation. Occasionally, Z as defined in Eq. (11) is used to evaluate the efficiency of H2O2 usage in the Fenton process (Kang and Hwang, 2000): Z ¼ 2:12CODoxi =½H2 O2 

(11)

In Eq. (11), CODoxi indicates the COD removed by oxidation, and [H2O2] indicates the dosage of added peroxide. At a low H2O2 concentration relative to the COD concentration in the untreated leachate, Z may be above 100% (Kim et al., 1997; Lau

et al., 2001). The observation suggests that in these cases the organic compounds are primarily oxidized by OH radical (formed from H2O2, or produced by photolysis in the photoFenton process) instead of by H2O2 because hydroxyl radicals may react with organic compounds to generate organic radicals that react immediately with dissolved oxygen (DO) to yield peroxyl radicals, thus initiating subsequent oxidation chain reactions. However, Z decreases gradually below 100% with increasing [H2O2], indicating that remaining organics are refractory and unfavorably oxidized (Kang and Hwang, 2000), or that some of the dOH produced is scavenged by excess H2O2 through Eq. (3). On the other hand, the amount of iron salt is also important to maximize net production of dOH, because excess iron salt also scavenges dOH through Eq. (4). An optimal molar ratio of H2O2 to iron salt required to maximize production of dOH has been theoretically calculated for some individual organic compounds. For example, Tang and Huang (1997) concluded that the optimal molar ratio of H2O2 to Fe2+ is 11:1, for Fenton oxidation of TCE. Such calculation is not possible for leachate, however, due to its chemical complexity and because a high dosage of iron salt is desired to promote coagulation and flocculation. Roddy and Choi (1999) reported that molar ratios of H2O2 to Fe2+ for batch mode and continuous mode operation were 1.5:1 and 3:1, respectively. Kim and Huh (1997) found an optimal molar ratio of 12.5:1 in batch tests. And, Lopez et al. (2004) reported a molar ratio of H2O2 to Fe2+ of almost 20:1, for treatment of a raw leachate in batch tests. The discrepancy in these results reflects variability in landfill leachate characteristics. The photo-Fenton method can be used to enhance reduction of Fe3+ to Fe2, and initiate production of dOH through photolysis, so that less Fe2+ is required. As a result, a higher ratio of H2O2 to Fe2+ is employed and less sludge is produced during photo-Fenton treatment of leachate. Kim and Vogelpohl (1998) reported ratios of H2O2 to Fe2+ of 26.2:1 and 35.0:1 for photo-Fenton and photo-Fenton-like processes, respectively, much larger than those reported for conventional Fenton treatment of leachate treatment.

7.3.

DO

Effect of DO for Fenton treatment of leachate is controversial in previous investigations. During Fenton oxidation DO drops rapidly to near zero in the presence of organic compounds, because the rate of O2 transfer from the atmosphere to solution is significantly lower than the O2 consumption rate. Kim and Vogelpohl (1998) found that air-saturated conditions strongly improved the reduction of TOC in photo-Fenton treatment of a biological pretreated leachate, compared with N2-saturated conditions. Rivas et al. (2003b) found that COD in an oxygen saturated solution dropped more rapidly than in a helium-saturated solution, in Fenton-like treatment of fermentation table olive brines having an initial COD of 15,000 mg/L. However, when they treated a landfill leachate using aerated and nitrogenated Fenton-like reactors, the effect of maintaining higher DO was not observed (Rivas et al., 2003a). Utset et al. (2000) investigated replacement of H2O2 by O2 during degradation of aniline by Fenton and photoFenton processes. They found that O2 only partially replaced the role of H2O2 in TOC reduction, and that the amount of

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4 0 (200 6) 368 3 – 369 4

O2 replacing H2O2 as an oxidizing agent depended on the concentrations of aniline and H2O2, pH, temperature, O2 flow rate and O2 feeding type, regardless of ferrous concentration.

7.4.

Effluent pH

Effluent pH subsequent to Fenton oxidation is typically adjusted to satisfy requirements for discharge or following biological treatment, and to convert dissolved iron to iron sludge. Although Fe3+ is least soluble at pH 8.0 (Stumm and Morgan, 1996), Rivas et al. (2004) found that pH 5.5 was most favorable for coagulation of effluent from Fenton oxidation of a leachate pretreated by coagulation. And, Kang and Hwang (2000) investigated operating pH ranging from 2.0 to 9.0 for coagulation of a leachate after Fenton oxidation, and found that both overall COD removal efficiency and COD removal efficiency by coagulation peaked at pH 3.0–6.0. These pH ranges are similar to the optimal pH reported in previous studies of single ferric salt coagulation for landfill leachate treatment (Christensen et al., 1992; Gau and Chang, 1996).

7.5. Addition modes of reagents and multiple-stage Fenton treatment Modification of the mode of addition of Fenton reagents, and two-stage Fenton treatment, have been reported for improvement of the removal of recalcitrant organics in leachate. The two similar operational modes involve successive addition of Fenton reagents to leachate. However, stepwise addition implies constant dosages of Fenton reagents to one unit, whereas two-stage treatment implies different dosages in two separate Fenton units. Generally, Fenton reagents are added to leachate in a single-step. However, such addition may cause self-decomposition of H2O2 due to high-localized concentrations at the point of injection, and scavenging of hydroxyl radicals by a large amount of hydrogen peroxide. To ensure complete usage of H2O2 for oxidation, several investigators added Fenton reagents in successive steps, but found different effects of stepwise addition. Yoo et al. (2001), Zhang et al. (2005), Deng (2006), and Englehardt et al. (2006) reported that stepwise addition of Fenton reagents increased COD removal, and COD removal augment increased with increasing addition times. However, Rivas et al. (2003a) found that COD removal efficiencies were not greatly different for single, two, and three-step addition of Fenton-like reagents to leachate. The discrepancy in these findings is unclear. The two-stage Fenton method for leachate treatment was used by Gau and Chang (1996) to treat a biologically pretreated leachate. COD was reduced from 2130 to 640 mg/L in the first Fenton stage using 200 mg/LH2O2 and 800 mg/LFe2+, and further decreased to 320 mg/L using 1000 mg/LH2O2 and 500 mg/LFe2+. Thus, two-stage Fenton treatment improved COD removal while consuming more reagents in the second stage, perhaps due to the recalcitrance of the organics remaining after first-stage treatment.

7.6.

Recycling of iron sludge

Recycling of iron sludge from the Fenton process to a coagulation step prior to Fenton treatment (Fig. 1(e)) can

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increase COD removal, decrease coagulant consumption, increase the settling velocity of coagulated sludge, and reduce the production of sludge. Yoo et al. (2001) carried out a series of lab and pilot-scale experiments to investigate the effect of Fenton sludge recycling. They found that addition of Fenton sludge did not impart an additional COD loading to the coagulation process and, on the contrary, the Fenton sludge could be used as an alternative coagulant to enhance COD removal efficiency and reduce the coagulant (ferric salt) dose up to 50%. In addition, sludge settling velocity was increased and less sludge production was observed.

7.7.

Temperature

Temperature has a positive effect on organic removal in Fenton treatment of landfill leachate. Rivas et al. (2003a) and Zhang et al. (2005) investigated 10–30 1C and 13–37 1C, respectively, and found that final COD removal increased with increasing temperature. However, a further temperature increase may significantly cause inefficient H2O2 decomposition that offsets increase of COD removal. As a result, the increase of COD removal is marginal in a high temperature. It should be noticed that the increase of organic removal due to temperature increase is relatively small compared to due to other factors (Zhang et al., 2005).

7.8.

UV irradiation

In the photo-Fenton process, UV irradiation is used to enhance the reduction of Fe3+ to Fe2+ and the resulting Fenton reaction, and the photolysis of H2O2 directly to OH. Kim et al. (1997) evaluated the effect of UV irradiation on photo-Fenton treatment of a mature leachate. They found that a specific energy input of 80 kW/m3 resulted in 51% TOC removal, three times the removal efficiency of the dark process. A specific energy input of 160 kW/m3 promoted higher TOC reduction. TOC reduction kinetics was described by pesudo-first order kinetics, with rate constants at 0, 80 and 160 kW/m3 UV input of 1.7, 9.6, and 18.7/min, respectively.

8.

Conclusion

As indicated in this survey, the Fenton reaction can be effectively exploited to treat landfill leachate, and may be particularly appropriate for mature leachate. Fenton processes are applicable to the treatment of highly toxic leachate and exhibit noticeably faster kinetics as compared with biological treatment processes. Fenton treatment also achieves considerably higher efficiency of removal of organic compounds from leachate, compared with other physical/ chemical technologies including coagulation and activated carbon adsorption. In particular, Fenton treatment effects chemical destruction rather than pollutant transfer from one phase to another, in contrast with membrane filtration and with other separation processes. Importantly, capital costs of Fenton treatment can be expected to be very low compared with other advanced oxidation processes such as UV/H2O2 (Tang, 2003). Also, energy requirements are low for conventional Fenton treatment, and operating costs depend

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primarily on the reagent concentrations. And Rivas et al. (2003a) made a rough estimation of operational costs for Fenton treatment of leachate, and obtained final cost of 8.0  103 US$ per m3 of leachate and ppm of COD removed, thus indicating that Fenton treatment may represent an attractive economical option. Furthermore, Fenton method for treatment of leachate was compared with other oxidation technologies such as ozonation, UV irradiation, H2O2 oxidation alone, ozone combined with UV or H2O2 by Englehardt et al. (2006). Fenton treatment shows great attraction in both removal efficiency of organic compounds and operating cost. Several drawbacks should be noted that may limit the wide application of Fenton treatment of leachate. First, as mentioned previously, some organic compounds are recalcitrant to Fenton treatment. Second, very significant quantities of acid are required to adjust the pH of typical leachate, significantly increasing the TDS of the effluent and presenting operational hazards and safety and corrosion issues. Third, sludge disposal after treatment is required, and the amount of sludge generated may influence operational costs. Finally, at temperatures less than 18.3 1C, the Fenton reaction may exhibit slow initial kinetics. Fenton treatment can be implemented alone or in tandem with other technologies for leachate treatment. Coagulation and oxidation both contribute to the removal of pollutants from leachate, and their relative importance depends on leachate characteristics and reaction conditions. Generally, an initial pH between 2.0 and 4.5 favors the Fenton reaction. In addition, the ratio of Fenton reagents greatly influences process efficiency because it determines the degree of scavenging of hydroxyl radicals. Moreover, aeration, final pH, reagent addition mode, temperature and UV irradiation may influence the efficiency of treatment of landfill leachate. When optimized accordingly, Fenton treatment can reduce COD, increase the ratio of BOD5/COD, and remove color in landfill leachate. R E F E R E N C E S

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