Abstract
In the past few years, there have been many researches on the use of different types of homogenous catalyst for the degradation of textile wastewater in conventional advanced oxidation processes (AOPs). However, homogenous AOPs suffer from few limitations, including large consumption of chemicals, acidic pH, high cost of hydrogen peroxide, generation of iron sludge, and necessity of post-treatment. Therefore, recently, there have been more researches that focus on improving the performance of conventional AOPs using heterogeneous catalysts such as titanium dioxide, nanomaterials, metal oxides, zeolite, hematite, goethite, magnetite, and activated carbon (AC). Besides, different supports such as AC that have been incorporated with transition metals and clays have been proven to have excellent catalytic activity in AOPs. This paper presents a comprehensive review of advances and prospects of catalytic AOPs for the decontamination of a wide range of synthetic and real textile wastewater. This review provides an up-to-date critical review of the information on the degradation of various textile dyes by a wide range of heterogeneous catalysts and adsorbents. The future challenges of AOPs, including chemical consumption, toxicity assessment, reactor design, and limitation of catalysts, are discussed in this paper. In addition, this paper also discusses the presence of ions, generation of by-products, and industrial applications of AOPs. Special emphasis is given to recent studies and large-scale combination of AOPs for wastewater treatment. This review paper concludes that more studies are needed for the kinetics, reactor design, and modeling of hybrid AOPs and the production of their corresponding intermediate products and secondary pollutants. A better economic model should also be developed to predict the cost of AOPs, as the treatment cost varies with dyes and textile effluents.
- Abbreviations
- AOPs
advanced oxidation processes
- BOD
biological oxygen demand
- COD
chemical oxygen demand
- CR
color removal
- DOC
dissolved organic compounds
- Fe2+
ferrous ion
- Fe3+
ferric ion
- H2O2
hydrogen peroxide
- ·OH
hydroxyl radical
- RT
retention time
- T
temperature
- TOC
total organic carbon
1 Introduction
Industrial development produces a large amount of wastewater that is extremely harmful to humans and the environment (Trabelsi-Souissi et al. 2011). Textile industry is one of the industries that consume a large amount of water (H2O), fuel, and chemicals. Effluent from the textile industry usually contains a high concentration of dyes and a variety of recalcitrant organic compounds. The characteristics of textile wastewater depend on the raw materials used, the processes involved, and various other factors. Synthetic dyes are extensively used for dyeing and printing in textile industries. There are more than 10,000 dyes with an annual production of more than 7×105 metric tonnes that are commercially available worldwide (Engin et al. 2008). It is estimated that approximately 15% of the dyestuffs are lost in industrial effluents during manufacturing processes (Ghodbane and Hamdaoui 2009). Such estimation indicates that 100 L of wastewater are produced for every kilogram of textile product, which is equivalent to 3.7 million liters each day worldwide (Amaral et al. 2014). The release of this colored effluent into environment affects the aesthetic properties of aquatic environments, interferes with the growth of bacteria, and causes chronic health effects to living organisms due to their toxicity and mutagenicity (Gunukula and Tittlebaum 2001).
Conventional wastewater treatment technologies often fail to make sure that treated effluent quality meets the national standards due to the increasing complexity of textile effluents. Therefore, a more efficient wastewater treatment system is needed to overcome the existing challenges (Tony et al. 2012). Various treatment technologies have been proposed, tested, and applied to meet both the current and the anticipated treatment requirements. The treatment processes must assure the elimination or mineralization of pollutants to meet the national standards of effluent quality (Hammami et al. 2012).
Chemical oxidation process is one of the solutions, and it can be divided into two classes: classical chemical treatments and advanced oxidation processes (AOPs; Oller et al. 2011). In the past 30 years, many studies have been conducted to investigate the efficiency of AOPs in treating different recalcitrant wastewaters that contain refractory and toxic pollutants (Balcıoğlu et al. 2003). AOPs are based on the generation of hydroxyl free radicals, which have a high electrochemical oxidant potential (Gaya and Abdullah 2008, Soon and Hameed 2011, Nezamzadeh-Ejhieh and Amiri 2013, Nezamzadeh-Ejhieh and Shirvani 2013, Nezamzadeh-Ejhieh and Khodabakhshi-Chermahini 2014). Having the strongest oxidizing power, they readily attack or degrade almost all recalcitrant organic compounds to carbon dioxide (CO2), H2O, and inorganic ions via hydroxylation or dehydrogenation. AOPs are widely used in wastewater treatment for overall organic content reduction, specific pollutant destruction, sludge treatment, and color and odor reduction (Gunukula and Tittlebaum 2001, Kavitha and Palanivelu 2004, Lucas and Peres 2006, Nezamzadeh-Ejhieh and Banan 2013). The feasibility of AOPs for wastewater treatment is further enhanced by their ability to produce hydroxyl radical (·OH) through different means.
The most widely researched AOPs include heterogeneous photocatalytic oxidation (Bauer et al. 1993), ozone (O3)-based processes (Meric et al. 2005), Fenton reaction, photo-Fenton reaction (Bouafia-Chergui et al. 2010, Oturan et al. 2011, Borba et al. 2013), and electrochemical oxidation. These techniques are promising in treating contaminated groundwater, surface H2O, and wastewater containing nonbiodegradable organic pollutants. Based on the literature review, a number of combinations of different AOPs have been found to be more efficient compared to individual oxidation process due to the generation of more radicals and high energy efficiency. Besides, the use of homogenous and heterogeneous catalysts and energy-dissipating components significantly enhances the degradation of textile effluent. Gogate and Pandit (2004b) presented an overview of hybrid methods, such as ultrasound (US)/hydrogen peroxide (H2O2) or O3, ultraviolet (UV)/H2O2 or O3, O3/H2O2, sonophotochemical oxidation, and photo-Fenton processes. Besides, Gaya and Abdullah (2008) presented a review on the fundamentals of the heterogeneous photocatalytic degradation of organic contaminants and reported the advantages of using titanium as a catalyst. There are very few studies that provide a comparative review on different types of homogenous and heterogeneous catalytic systems for textile wastewater treatment. Therefore, this work focuses on reviewing the limitations of different types of conventional homogenous AOPs and the feasibility of using various heterogeneous catalysts for textile wastewater treatment to enhance the treatment efficiency.
This review aims to highlight the most recent advances and prospects of catalytic AOPs for textile wastewater treatment. This article describes the composition of textile wastewaters and the limitations of the current conventional methods to effectively degrade textile effluent. In addition, this review also presents various applications of homogenous AOPs for textile wastewater treatment, the mechanism of AOP reactions, and the key parameters that affect the processes. The limitations of conventional AOPs and the solutions to overcome those challenges are also suggested in this paper. It is crucial to identify and develop the most effective treatment method that is economically feasible and eco-friendly. Therefore, this review paper provides a comprehensive information on the degradation of various textile dyes by a wide range of heterogeneous catalysts and adsorbents. Besides, the authors also discussed the limitations of conventional AOPs for industrial application and suggested solutions to overcome the limitations.
2 Textile wastewater
Textile dyeing and finishing processes consume a significant amount of H2O and chemicals, generating a remarkably large amount of wastewater and making it the most polluting industrial effluent (Dantas et al. 2006). Textile processing involves many different processes and the composition of wastewater depends greatly on the processes and chemicals used. The general textile processing steps include sizing and desizing, scouring, bleaching, dyeing, and finishing (Vilar et al. 2011). Scouring, bleaching, and dyeing generate a large amount of wastewater, which varies in composition, whereas sizing produces little but concentrated wastewater (Hassan et al. 2009).
The pollutants that are present in textile wastewater in large quantity are unfixed dyes or dispersing agents that are organic and highly colored, which are washed out during the dyeing process. Textile wastewater can cause severe impacts on the environment by obstructing sunlight penetration into aquatic environment, reducing reoxygenation, affecting the aesthetic value of H2O resources, and interfering with aquatic biological processes that eventually affect the ecobalance (Torrades et al. 2004). Some major pollutants in textile effluent originate from the synthetic dyes that are extensively produced and used as a colorant in textile, leather, painting, and printing industries in the 21st century due to their unique properties, such as high wet fastness profile, brilliant shades, low cost, and simple synthesis methods. It is estimated that more than 100,000 commercial dyes with an annual production of 7×105 tonnes are available and most of them are synthetic and soluble compounds.
2.1 Classification of dyes
Synthetic dyes are categorized as acidic, reactive, disperse, vat, metal complex, mordant, direct, basic, and sulfur dyes based on their application (Reife and Freeman 1996). Table 1 shows the classification of dyes based on their applications.
Classification of dyes based on the application.
| Dyes | Application | Chemicals |
|---|---|---|
| Reactive | Wool, cotton, silk, and nylon | Anthraquinone, formazan, phthalocyanine, azo, oxazine, and basic |
| Acid | Wool, nylon, and silk | Anthraquinone, xanthene, azo, nitro, and triphenylmethane |
| Basic | Nylon and polyester | Hemicyanine, azo, cyanine, diazahemicyanine, azine diphenylmethane, xanthene, triarylmethane, acridine, anthraquinone, and oxazine |
| Direct | Nylon, rayon, and cotton | Phthalocyanine, azo, oxazine, and stilbene |
| Disperse | Polyamide, acrylic polyester | Benzodifuranone, azo, anthraquinone, nitro, and styryl |
| Sulfur | Rayon and cotton | Indeterminate structures |
| Vat | Wool and cotton | Indigoids and anthraquinone |
Synthetic dyes that are frequently used at industrial scale contain complex aromatic amide groups with alkyl, halogen, nitro, hydroxyl, sulfonic acid, substituent(s), and inorganic sodium salts (Technology of Synthetic Dyes, Pigments and Intermediates 2005). Besides, industrial dyes are generally toxic, carcinogenic, and stable because they contain azo, anthraquinone, sulfur, triphenylmethyl, and phatalocianine groups in their structure. Therefore, dye effluents contain a high concentration of recalcitrant compounds, toxic substances, detergents and soaps, oil and grease, sulfides, soda, and alkaline-rich wastes. It is estimated that approximately 12% to 15% of the dyes are lost and discharged in the effluent during the wet processing activities (Al-Amrani et al. 2014), because dyes are not completely fixed to the fiber during the dyeing process. Figure 1 gives the percentage of unfixed dyes for various dyes together with their applications. Generally, reactive dyes have the poorest fixation.
Percentage of unreacted dyes from industrial process based on the type of dyes.
2.2 Characterization of textile wastewaters
Recalcitrant textile wastewater typically contains high concentrations of organic compounds [expressed as high chemical oxygen demand (COD)], colors associated with residual dyes, a wide range of pH, low biodegradability, and high salt content (Balcioglu and Arslan 1997, Lahkimi et al. 2006, Schrank et al. 2007, Hassan et al. 2009, Rodrigues et al. 2009, Blanco et al. 2012, Hammami et al. 2012, Punzi et al. 2012). The characteristics of textile wastewater are depicted in Table 2.
Typical textile industrial wastewater characteristics.
| Parameters | Range |
|---|---|
| COD | 150–12,000 mg/l |
| Total suspended solids | 2900 and 3100 mg/l |
| Total nitrogen | 70–80 mg/l |
| BOD | 80–6000 mg/l |
| BOD/COD | 0.25 |
BOD, Biological oxygen demand.
The H2O environment can be adversely affected by just a small amount of dyes. Therefore, the removal of organic contaminates from wastewater has become an environmental concern. It is more complicated to treat colored industrial effluents because they are highly inconsistent in both the hues and the concentrations of color.
Conventional techniques based on physicochemical, chemical, and microbiological methods, such as coagulation, adsorption, ion exchange, and ultrafiltration, are traditionally used for degrading persistent organic pollutants that can be found in textile wastewater. Although these methods yield significant color removal (CR), they are unable to ensure that the treated effluent meets the discharge criteria of wastewater in terms of COD. These methods also generate solid wastes that require further handling, as they only transform pollutants from one phase to another but do not degrade them (Wu et al. 2008a, Punzi et al. 2012, Babuponnusami and Muthukumar 2013). In addition, adsorbent regeneration, excess sludge production, and rapid fouling of the used membranes are some of the other disadvantages of conventional treatment processes. The characteristics of persistent organic pollutants, such as high stability to sunlight irradiation, resistance to microbial attack, and temperature, impair their removal efficiency by conventional wastewater treatment.
It has been frequently discussed that conventional treatment methods such as biological and simple physicochemical treatments are not efficient for complete mineralization, as most of the pollutants are highly recalcitrant (Bandala et al. 2008, Karthikeyan et al. 2011).
New environmental concerns and regulations are urging the textile industry to reduce pollutants and recycle the process H2O, as textile effluent rarely meets the discharge standards (Karthikeyan et al. 2011). Based on the review conducted on the limitations of the methods, it is necessary to conduct an intensive research to identify more effective wastewater treatment systems. The treatment should focus on the removal of biodegradable components of wastewater followed by the degradation of recalcitrant contaminants. Alternative treatment technologies capable of mineralizing refractory molecules effectively, such as pesticides, surfactants, coloring matter, and endocrine-disrupting chemicals, are highly needed. Therefore, developing a practical and efficient H2O treatment system has been of global interest in the past decades. AOPs, based on the in situ generation of powerful oxidizing agents such as ·OH, have therefore been studied by many researchers in the last two decades. AOP is a complete, irreversible degradation process that does not produce toxic and instable products.
3 Advanced oxidation processes
The concept of “advanced oxidation technologies” was established by Glaze et al. (1987). AOP is a treatment process that involves the generation of highly reactive ·OH. This technology has gained popularity as an imperative technology for the oxidation and destruction of refractory compounds in wastewater (Tony et al. 2012). Oxidizing power is one of the most important factors to be considered when choosing an oxidant for oxidation process. ·OH is the second strongest oxidizing species with a relative oxidation power of 2.8 eV. This nonselective primary oxidant is capable of readily attacking or degrading almost all recalcitrant organic compounds and converting them to less harmful intermediate products (Vujević et al. 2010). ·OH attack organic molecules through three possible means: (i) dehydrogenation of a hydrogen atom to form H2O, (ii) hydroxylation of a nonsaturated bond, and (iii) redox reaction. Figure 2 illustrates some special characteristics of ·OH that make AOPs a powerful method for the removal of refractory compounds.
Features of ·OH.
AOPs can be further classified based on the existing literature information. Table 3 summarizes the means of hydroxyl production by AOP. The feasibility of AOP for wastewater treatment is enhanced by the fact that they offer multiple means for ·OH generation. Literatures show that ·OH with a high oxidizing potential (Tony et al. 2012) can be generated through ozonation (Gozzi et al. 2012), Fenton process, Fenton-like process (Prousek et al. 2007, Duarte and Madeira 2010, Wasewar et al. 2011), photochemical oxidation, electrochemical oxidation (Jain et al. 2011, Oturan et al. 2011, Babuponnusami and Muthukumar 2012), H2O2/UV, UV/O3 (Lucas and Peres 2007, Schrank et al. 2007), photocatalysis oxidation (Bauer et al. 1993, Karthikeyan et al. 2011), and sonolysis (Zheng et al. 2007).
AOPs classification based on the mechanism.
| Reaction | Mechanism | AOPs |
|---|---|---|
| Homogeneous | Chemical | O3 |
| O3/H2O2 | ||
| Fenton | ||
| Electrochemical | EF | |
| PEF | ||
| Photochemical | O3/UV | |
| O3/H2O2/UV | ||
| H2O2/UV | ||
| Photo-Fenton | ||
| Heterogeneous | Photochemical | TiO2/O2/UV |
| TiO2/H2O2/UV |
AOP-based methods are promising for the abatement of numerous organic pollutants in wastewater. This technique is promising in converting toxic and biologically resistant compounds into less harmful compounds (Tekin et al. 2006). The successful implementation of AOPs highly depends on necessary pretreatments that maximize the reduction of suspended solids, colloids, and grease.
3.1 O3-based AOPs
Ozonation has been used as the oxidation technology for textile wastewater by many researchers in the last two decades. AOPs of wastewater and the use of O3 are two-phase gas-liquid reaction. Ozonation always involves these two species: O3 and ·OH. The chemical reaction of aqueous O3 involves the decomposition of O3 via a chain reaction mechanism and results in the generation of ·OH. One of the advantages of ozonation is that O3 can decompose into ·OH. At low pH, molecular O3 is predominant, where organic compounds are subjected to electrophilic attraction and decomposition. On the contrary, at higher pH, O3 molecules decompose into free radicals (·O2- and HO2·) and subsequently two ·OH. These radicals are free to react with organic pollutant at rapid rates (rate constant=108–1010 M-1 s-1) (Cooper and Burch 1999). O3 might react directly with organic compounds at rates varying by several orders of magnitude. O3 attacks the organic contaminants in wastewater through direct reaction (as molecular O3) or by forming free ·OH via a complex decomposition process (Legube et al. 1985, Wang and Xu 2012). It should be noted that O3 is a powerful oxidizing agent (E0=2.07 V) with a high oxidation potential. As a result, it can effectively break down the aromatic compounds in wastewater and conjugate double bonds in dye chromophores (Ikehata and El-Din 2004).
Many researchers have reported the efficiency of ozonation for disinfection, decolorization, dewatering, degradation of toxic synthetic organic compound traces, and removal of persistent organic pollutants, odor, sludge, and foam (Amat et al. 2005, Gong et al. 2008, Wu et al. 2008d, Bustos et al. 2010, Biń and Sobera-Madej 2012, Zhao et al. 2013, Ding et al. 2014). Ozonation has been found effective for the decolorization and mineralization of synthetic dyes. The conditions that contribute to the formation of O3 and ·OH vary with the characteristics of pollutants, H2O quality, concentrations of pollutant, temperature, pH, and level of ozonation. Figure 3 shows the ranges of O3 dosages required regardless of its pollutant type.
Ranges of O3 dosages regardless its treatment (adapted from Ried et al. 2009).
Ozonation has several advantages: (i) it does not increase the volume of wastewater, (ii) it does not produce sludge, (iii) it removes color and organic matter (OM) in one step, (iv) it needs only a little space, (v) there is easy on-site installation, (vi) it is less harmful than other oxidative processes because no stock H2O2 or other chemicals are required on-site, and (vii) residual O3 can be easily decomposed to oxygen (Alvares et al. 2001, Oguz and Keskinler 2008, Chen et al. 2009b, Loeb et al. 2012).
On the contrary, there are several disadvantages associated with ozonation, which limit its application in larger- and industrial-scale textile wastewater treatment. The main disadvantage of ozonation is high consumption of energy because it uses electrical energy. Besides, O3 is unstable and it requires complex mixing techniques that are associated with high equipment and maintenance cost. In addition, there is also potential bromate formation.
3.1.1 O3 combined with H2O2
Researchers have discovered that the limitations of ozonation can be overcome by combining it with other oxidants. O3 decomposition in aqueous solution is achieved through the formation of ·OH. ·OH acts as an initiator in the reaction mechanism. The O3/H2O2 process is recognized as an effective chemical oxidation method to mineralize recalcitrant organics in wastewater, as the addition of H2O2 enhances O3 decomposition with the formation of ·OH. The reaction rate constant of ·OH (106–109 M-1 s-1) is several orders higher than that of O3 (Spivey et al. 2005). H2O2 is an inexpensive, highly soluble, and readily available oxidizing agent that generates ·OH and initiates the decomposition of O3 when mixed with O3. The reaction between H2O2 and O3 takes place when it is present as an anion, HO2-. H2O2 reacts with O3 after it is partially dissociated into hydroperoxide anion, HO2-. Previous studies have shown that better results could be obtained while treating wastewater using the O3-H2O2 combination compared to O3 alone. The reaction is described below [Equations (1)–(4); Rosenfeldt et al. 2006]:
The combined reaction shows that two O3 molecules produce two ·OH.
3.1.2 O3 combined with UV radiation
Ozonation assisted by UV radiation uses UV photons to accelerate O3 molecules to produce strong ·OH (Glaze et al. 1987). Irradiation with UV was shown to reduce the required amount of O3 and inhibit the formation of bromate. It was also reported that the sequential application of O3 and UV significantly increased the degradation and biodegradability of the textile effluents. O3 reacts with UV radiation to form ·OH. The photodecomposition of O3 leads to the production of intermediate H2O2 (Rao and Chu 2010).
This system contains two components, which are UV radiation and O3, to oxidize pollutants. The UV lamps used in this system must have a maximum radiation output of 245 nm for efficient O3 photolysis. The most commonly used light source is medium-pressure mercury lamp, which generates UV light at a wavelength of approximately 280 nm.
3.1.3 O3 combined with H2O2 and UV radiation
The combination of O3, H2O2, and UV radiation mainly generates ·OH with high oxidation power compared to O3/H2O2 and O3/UV (Monteagudo et al. 2005). The system can be applied to highly polluted wastewaters for the effective degradation and mineralization of recalcitrant organic compounds. The advantages of the combined system are great reactivity and reduction-oxidation potential. ·OH react unselectively with recalcitrant organic components present in wastewaters (Katsoyiannis et al. 2011). This combined system also has a high degree of flexibility, where each of the treatment technique can be used individually or combined to oxidize different types of wastewaters. The system can also be applied at atmospheric ambient pressure and room temperature (Lucas et al. 2010). The oxidation power of H2O2 and O3 can be significantly enhanced with the use of UV radiation (Esplugas et al. 2002).
3.2 Fenton-based AOPs
3.2.1 Homogeneous Fenton process
Fenton process is capable of completely degrading contaminants and converting them into harmless compounds such as CO2, H2O, and inorganic salts (Rivas et al. 2003). This process generates ·OH through the reaction of H2O2 and ferrous salt in acidic environment. The use of Fenton processes for treating various types of wastewater has also been studied (Bauer et al. 1993, Perkowski et al. 2000, Modenes et al. 2012, Tony et al. 2012, Borba et al. 2013). Homogeneous Fenton process uses Fe salts as a catalyst and the readily available Fe ions in the reaction medium react effectively in the degradation process. It should be noted that mass transfer limitations are negligible in the classical Fenton process.
Ferrous ions (Fe2+) initiate and catalyze the H2O2 oxidation process to generate highly reactive short-lived ·OH (Lucas and Peres 2007), as depicted in Equation (8) (Rigg et al. 1954, Walling 1975, Buxton et al. 1988, Neyens and Baeyens 2003). The generation of the radicals involves a complex reaction sequence in an aqueous solution.
Furthermore, the ferric ions (Fe3+) formed can react with H2O2, causing it to regenerate Fe2+ and form ·OH and hydroperoxyl radical during the process. The reactions are shown in Equations (10) to (14) (Haber and Weiss 1934, Walling and Goosen 1973, Pignatello et al. 2006, Jung et al. 2009, Garrido-Ramírez et al. 2010):
The reaction of H2O2 with Fe3+ is referred to as a Fenton-like reaction and shown in Reactions (10) and (11) (Walling and Goosen 1973, Bielski et al. 1985).
Based on Reaction (14), H2O2 can act as a ·OH scavenger as well as an initiator [Reaction (8)]. ·OH can oxidize organics (RH) by the abstraction of protons producing organic radicals (R·), which are highly reactive and can be further oxidized (Walling and Kato 1971).
Fenton oxidation has been studied extensively by researchers, as it produces high chemical activities. It is also cost-efficient and easily obtainable. ·OH produced via Fenton reaction can remove or degrade a high amount of various recalcitrant organic compounds (Hu et al. 2011). There has been a remarkable development on Fenton application for H2O treatment.
Fenton process appears to have the capability to completely decolorize and mineralize wastewater from different sources, such as leachate, textile dyeing, pharmaceutical, petroleum, olive oil processing, paper mill, and tannery industries within a short period of time (Lopez et al. 2004, Lucas and Peres 2006, 2009, Shemer et al. 2006, Liu et al. 2007). It is able to completely destroy the chromophoric structure of azo dyes and the degradation competency depends on the structure and nature of auxiliary groups attached to the aromatic nuclei of the dyes. In the study by Singh and Tang (2013), total COD removal of 30% to 95% for landfill leachate with initial COD concentration ranging from 93,000 to 94,920 mg/l was achieved through Fenton process. High COD removal was observed because ·OH generated in Fenton process could effectively transform low-molecular-weight biodegradable OM into more biodegradable matter that can be subsequently treated using biological processes.
Some of the factors influencing the applicability of Fenton process for real industrial wastewater include high chemical consumption (up to 50–80 ppm of Fe ions in solution), pH dependence (effective in the range of pH 2–4), high cost of H2O2, generation of Fe sludge, necessity of post-treatment, necessity of neutralization of the treated sample before disposal, and decomposition of H2O2 (Iurascu et al. 2009, Babuponnusami and Muthukumar 2012, Ali et al. 2013). Besides, the treatment of Fe ion-containing sludge as the last wastewater treatment step is expensive and it requires a huge amount of chemicals and manpower.
The practicality of Fenton process is severely restricted by all the above-stated limitations. Therefore, there is a need to modify the conventional Fenton oxidation process to improve the oxidation efficiency and reduce the operating cost by reducing the reagent usage. There has been continuous research on exploring the alternatives to modify the classical Fenton process.
Great efforts have been made to develop Fenton oxidation based on the heterogeneous catalytic system. Heterogeneous Fenton oxidation is cost effective in overcoming the shortcomings of homogenous Fenton oxidation and it can also work in a wide range of pH. It reduces the loss of catalysts, yields a high reaction rate, encourages the dispersion of Fe oxides (FeO), and generates very little Fe sludge. The use of goethite (α-FeOOH), hematite (α-Fe2O3) clay minerals, Fe hydroxide, and Fe supported on other materials, such as silica and alumina, to replace soluble Fe salts has been studied.
3.2.2 Homogeneous photo-Fenton process
Fenton process can be enhanced by UV irradiation or UV-visible light. The complete mineralization of numerous organic compounds in aqueous solutions can be achieved through the application of UV irradiation. The efficiency of UV irradiation is attributed to the photoreduction of Fe3+ to Fe2+ ions, which produces new ·OH with H2O2 (Navarro et al. 2010).
Photo-Fenton reaction is based on the generation of ·OH produced by the reaction between H2O2 and Fe2+ as a catalyst under UV-visible light irradiation (Kortangsakul and Hunsom 2010). Fe3+ complexes, [Fe3+(OH)-]+ and [Fe3+(RCO2)-]2+, are the main light-absorbing compounds in the system, which produce additional Fe2+ through photo-induced, ligand-to-metal charge transfer reactions. Photo-Fenton oxidation is suitable for treating aromatic pollutants due to its effectiveness in degrading different types of organic compounds (Giroto et al. 2008, Lopez-Alvarez et al. 2012). It has been proven that photo-Fenton and Fenton-like processes enhance the reaction rate through the generation of high-valence Fe intermediates that directly attack the organic compounds. High-valence Fe-based oxidants are formed through the absorption of visible light by the complex formed between Fe3+ and H2O2. The degradation efficiency of photo-Fenton is strongly affected by solution pH, initial H2O2 and Fe2+ concentration, UV dosage, scavengers, temperature, and contact duration. The optimum pH for irradiation-assisted Fenton ranges from 2 to 5. Similar to Fenton process, at pH values higher than 5, Fe2+ ions are unstable and forms Fe3+, which may form complexes with ·OH.
Homogeneous photo-Fenton and photo-Fenton-like processes have been widely reported for the treatment of colored wastewater, as they generate strong ·OH and destroy organic pollutants significantly (Wu et al. 1999, Elmorsi et al. 2010, Archina et al. 2015). However, the cost associated with the processing of sludge and the narrow pH range of the treatment system has limited the application of homogeneous photo-Fenton systems for wastewater treatment (Pignatello et al. 2006, De la Cruz et al. 2012). Therefore, much attention has recently been paid to the development of heterogeneous catalysts to reuse catalysts and avoid the formation of metal-containing sludge that cause pollutions and requires post-treatment in photo-Fenton processes (Zhang et al. 2007b, Tekbaş et al. 2008, Chen et al. 2009a, An et al. 2013).
3.3 Heterogeneous photocatalytic oxidation
Among AOPs, heterogeneous photocatalysis appears as a promising solution in wastewater treatment (Balcioglu and Arslan 1997, Nezamzadeh-Ejhieh and Karimi-Shamsabadi 2014). This nonselective treatment degrades pollutants in the presence of catalysts as a semiconductor material and UV radiation (So et al. 2002). To date, researchers have used UV irradiation (wavelength ranging from 320 to 400 nm), visible light, and solar radiation in heterogeneous photocatalytic oxidation for the photodegradation of pollutants (Han et al. 2009, Nezamzadeh-Ejhieh and Khorsandi 2010a,b, Nezamzadeh-Ejhieh and Salimi 2011, Nezamzadeh-Ejhieh and Karimi-Shamsabadi 2013, Nezamzadeh-Ejhieh and Moazzeni 2013). When a photocatalyst is illuminated by radiation with an energy greater than the band gap of the semiconductor, valence band holes and conduction band electrons are generated. The valence band holes react with H2O and hydroxide ion to form ·OH, whereas electrons react with the molecular oxygen absorbed on the surface of the catalyst and form superoxide radical anions that, in turn, react with protons to form peroxide radicals (Nezamzadeh-Ejhieh and Hushmandrad 2010, Nezamzadeh-Ejhieh and Khorsandi 2010a,b, Nezamzadeh-Ejhieh and Moeinirad 2011, Zabihi-Mobarakeh and Nezamzadeh-Ejhieh 2015). It has been proven that the presence of catalysts enhances the rate of photodecomposition.
Titanium dioxide (TiO2), zinc oxide (ZnO), cesium oxide (CeO2), ZrO2, CdS, and ZnS are among the semiconductors used in a heterogeneous photocatalyst system for environmental treatment (So et al. 2002, Muruganandham and Swaminathan 2006b, Peternel et al. 2007, Gaya and Abdullah 2008, Asiltürk and Şener 2012, Sharma et al. 2012). Among them, ZnO and TiO2 have been widely investigated because they offer several advantages, such as being operable at ambient conditions, inexpensive, commercially available, nontoxic, and photochemically stable, compared to others. Besides, the morphology of TiO2 allows easy electron transfer and stabilizes charge separation. As a result, the recombination of photogenerated carriers is prevented. The addition of electron acceptors such as H2O2, potassium bromate (KBrO3), and ammonium persulfate into the solution increases the efficiency of photocatalytic oxidation by enhancing the generation of ·OH, reducing the chances of electron-hole recombination, increasing the oxidation rate of intermediate compounds, and diminishing problems caused by low oxygen concentration [refer to Reactions (16) and (17); Saquib and Muneer 2002, Nezamzadeh-Ejhieh and Moazzeni 2013]. At the same time, an excessive amount of H2O2 has a negative effect on the degradation and decolorization efficiencies. This is because the scavenging of ·OH occurs at higher H2O2 dosage and it produces a weaker oxidant, HO2· [Reactions (18) and (19); Chu and Wong 2004, Chen and Liu 2007, Nezamzadeh-Ejhieh and Salimi 2010]. This situation reduces the available ·OH for oxidation (Wang et al. 2010). Besides, an excess amount of H2O2 might also absorb and reduce the incident UV light used for the photocatalysis process (Chu and Wong 2004, Muruganandham and Swaminathan 2006a).
Some earlier studies have concluded that the extensive use of TiO2 is not economically viable for industrial-scale H2O treatment due to its wider band gap that can only be activated by UV light. However, latter investigation has shown that photocatalysts can be activated by solar radiation, which is a renewable energy. ZnS has attracted researchers in search for a more reliable photocatalyst, as it has nearly the same photodegradation mechanism as TiO2. However, ZnS has not been studied in detail due to some limitations. The major disadvantage of ZnS is that it can only work under UV irradiation due to its large band gap of 3.6 eV. On the contrary, the photocatalytic activity of ZnS can be improved by doping it with metals such as Cu, Ni, or Pb through coprecipitation method or polyreaction. Because doping is a complex process, developing a simple, economically viable, and environment-friendly method to synthesize ZnS is of concern.
3.4 Electrochemical oxidation
The recent industrial focus on sustainable development has drawn much attention from electrochemical technologies that use electrical energy. It is an environment-friendly technology due to its usage of electron, which is a clean reagent without the usage of chemical reagents for pollutant degradation. The commonly used electrochemical methods are electroreduction, electrodeposition, electrocoagulation, electroflotation, electrophotoxidation, electrodisinfection, and electrooxidation. Among electrochemical technologies, electrochemical oxidation has been widely researched due to its simplicity. Electrochemical oxidation works based on direct oxidation at the anode or indirect oxidation using suitable anodically formed oxidants.
Electrochemical AOPs (EAOPs) are known as promising AOPs for wastewater treatment because they are capable of treating a wide range of toxic and recalcitrant organic contaminants. It is based on the in situ production of ·OH, which plays a central role by reacting with organic contaminants and completely mineralizing them to CO2, H2O, and inorganic ions (Pipi et al. 2014). This technology is increasingly used by researchers for wastewater treatment because it is compact, simple, easy to operate, energy efficient, versatile, cost effective, amenable to automation, and environment friendly. Therefore, researchers have been studying EAOPs to find the most efficient material, operating parameters, and process to completely degrade highly polluted and recalcitrant wastewater. Several reviews have reported the application and prospects of EAOP for wastewater treatment (Gutiérrez and Crespi 1999, Martínez-Huitle and Brillas 2009, Brillas and Martínez-Huitle 2015, Fernandes et al. 2015).
The most efficient EAOP is based on the electrogeneration of H2O2 at cathodes from two-electron reduction of oxygen gas [Equation (20)].
Anodic oxidation and indirect electro-oxidation are the simplest EAOPs for the oxidation of recalcitrant compounds. The organic pollutants are commonly degraded by (i) direct electron transfer at the anode, which yields very poor decontamination, and (ii) ·OH that is formed as an intermediate product from H2O oxidation to O2 at the anode surface that leads to total or partial decontamination. Electrochemical conversion and electrochemical combustion are two approaches that have been proposed for electrochemical oxidation based on the existing indirect or mediated oxidation. Electrochemical conversion transforms the selected refractory organic compounds into carboxylic acid. Then, the complete mineralization of organic compounds to CO2 and inorganic ions is achieved through ·OH attack in electrochemical combustion methods.
Electro-Fenton (EF) and photo-EF (PEF) are two emerging electrochemical-based AOPs for wastewater treatment (Zhang et al. 2007a, Babuponnusami and Muthukumar 2013, Khataee et al. 2013a,b). In EF, H2O2 is directly generated at the cathode of the cell from O2 gas reduction. This indirect electro-oxidation method offers advantages over individual Fenton process such as on-site production of H2O2, higher degradation efficiency, continuous generation of Fe2+ at the cathode, and lower cost. PEF is a combination of electrochemical and photochemical processes with Fenton process. This combination generates more radical species, especially ·OH, compared to Fenton process. The radiation enhances the EF system via the photochemical regeneration of Fe2+ by the photoreduction of Fe3+.
EAOPs have attracted great attention due to the advantages offered by the treatment system such as high energy efficiency, versatility, and simplicity. In electrochemical oxidation, in situ H2O2 production and stability depend on cell configuration, types of cathode, anodic materials, and operating conditions. Therefore, it is important to optimize the operating parameters to achieve the highest degradation level with the lowest cost. The studies pertaining to the application of PEF process for dye degradation are very limited.
4 Key controlling parameters for the application of AOP technology
The performance efficiency of AOPs is greatly dependent on generation of ·OH. Various operating parameters that significantly affect the generation of ·OH have been described by many researchers. It is important to understand the relationship between operating parameters and ·OH production and consumption. Tables 4 and 5 reveal the relevant experimental work and operating parameters involved in AOPs for different types of H2O matrix. Based on the literature review, the parameters that highly affect AOPs include pH of the treatment system, type and initial concentration of catalysts and oxidants, temperature, presence of catalyst, retention time (RT), agitation speed, partial pressure of O3, light intensity, initial concentration of pollutant, matrix of pollutant, type of buffer used for pH adjustment, radiant flux, wavelength of irradiation, aeration, presence of ions species, reactor design and concentration of gas fed to the photoreactor (Ramirez et al. 2007a,b, Alvárez’ et al. 2009). The effects of all the above parameters on efficiency of AOP systems have been widely investigated (Gogate and Pandit 2004a, Alaton and Teksoy 2007, Lin et al. 2008, Bagal and Gogate 2014).
Homogeneous catalyst commonly used for dye degradation in AOPs.
| Method | Pollutant | Operating parameters | Performance | References |
|---|---|---|---|---|
| Fenton | ||||
| Fenton’s reagent using fluidized-bed reactor | RB5, reactive orange 16, and reactive blue 2 | [Dye]: 0.1 mm, [H2O2]: 0.16 g/l 74.07 g carrier/l, pH 3 | XCR: 96–99%, XCOD: 34–49% (100 min) | Su et al. (2011a) |
| Fenton | Real naphthalene dye | COD=7300 mg/l, [H2O2]=4.9 g/l, Fe2+/H2O2=1/20, pH 2.5 | XCOD: 93%, XTOC: 62% | Gu et al. (2012) |
| Fenton’s reagent-yeast | RB5 | [Dye]=500 mg/l, [H2O2]=1 mm, [Fe2+]=0.1 mm, pH 5, T=20°C | XCR: 91% (60 min) | Lucas et al. (2007) |
| Fluidized-bed Fenton process | Dyeing and finishing mill | [COD]=314–404 mg/l, [COD]:[Fe2+]:[H2O2]=1:0.95:3.17, pH 3, SiO2=74.07 g/l | XCR: 92%, XCOD: 87% (60 min) | Su et al. (2011b) |
| Fenton oxidation | Everdirect supra-turquoise blue FBL | [COD]=500 mg/l, [H2O2]=5.2, [Fe2+]=3.6, pH 2.45 | XTOC: 93.06% (15 min) | Karthikeyan et al. (2011) |
| Fenton oxidation | RB5 | [H2O2]0/[RB5]0=4.9:1, [H2O2]0/[Fe2+]0=9.6:1, pH 3.0 | XCR: 97.5%, XTOC: 21.6% (30 min) | Lucas and Peres (2006) |
| Fenton | C.I. Acid yellow 23 | [Dye]=40 mg/l [H2O2]=17mm, [Fe2+]=0.05 mm, pH 3 | XDye Removal: <60% (25 min) | Modirshahla et al. (2007) |
| Fenton | Peach red azo dye, HF6 | [Dye]=0.2 mm, [H2O2]=30 mm, pH 3, [Fe2+]=0.6 mm | K: 1.8×108 M-1 s-1 | Chang and Chern (2010) |
| Fenton | Textile wastewater | [COD]: 8100 mg/l, [H2O2]=0.882 mm, Fe2+=40 mg/l, pH 3 | XCR: 71.5%, XCOD: 45% | Papadopoulos et al. (2007) |
| Fenton biological treatment | RB5, reactive blue 13, and acid orange 7 | [Dye]=50 mg/l, [Fe2+]=0.27 mm, [H2O2]=1.47 mm, pH 3 | XCR: >95%, XOrganic Compound Reduction: 41–91% | Lodha and Chaudhari (2007) |
| Photo-Fenton | ||||
| Photo-Fenton | Carmoisine edible | [Dye]=20 mg/l, [Fe2+]=0.0125 mm, H2O2=0.3 mm, pH 3.5 | XDye Removal: 95.1% (30 min) | Sohrabi et al. (2014) |
| Photo-Fenton | Real dye wastewater | [H2O2]=73.5 mm, [Fe(II)]=1.79 mm, T=298 K, pH 3 | XCOD: 76.3% | Torrades and García-Montaño (2014) |
| Photo-Fenton | RB5 | [RB5]: 0.1 mm, [H2O2]0/[RB5]0=4.9:1, [H2O2]0/[Fe2+]0=9.6:1, pH 3.0 | XCR: 98.1%, XTOC: 46.4% (30 min) | Lucas and Peres (2006) |
| Sequencing batch reactor coupled to photo-Fenton processes | Real textile wastewater | [COD]=1560 mg/l, [Fe2+]=66.5 mg/l, [H2O2]=1518 mg/l, T=25°C, pH 2.7 | XCOD: 97%, XTOC: 95% (60 min) | Blanco et al. (2014) |
| Fenton/UV-C and ferrioxalate/H2O2/solar light processes | RB5 | [Dye]=0.1 mm [H2O2]=1.5 mm, [Fe2+]=0.5 mm, solar light intensity average=530 W/m2, pH 5 | XCR: 98.1%, XTOC: 46.4% (60 min) | Lucas and Peres (2007) |
| UV/Fenton | Mixed dyes | [Dye]=100 mg/l, Fe2+/H2O2=0.5 mm/20 mm, Fe0/H2O2=2 mm/1 mm, pH 3 | XCR: 78–84%, XTOC: 95–100% (90 min) | Papic et al. (2009) |
| Coupling a photo-Fenton system with a sequencing batch reactor | Procion red H-E7B | [Dye]=250 mg/l, [Fe2+]=0.29 mm, [H2O2]=3.7 mm, pH 3 | XTOC: 99% (60 min) | García-Montaño et al. (2006) |
| Photo-Fenton | Tartrazine (E102) | [Dye]=0.01 mm, [H2O2]=0.6 mm, [Fe]=0.08, pH 3 | XTOC: 80% (60 min) | Oancea and Meltzer (2013) |
| Photo-Fenton | Remazol red RR | [Dye]=100 mg/l, Fe2+/Fe0=0.25 mm, H2O2=12 mm, pH 3 | XTOC: 93% (60 min) | Punzi et al. (2012) |
| Photo-Fenton, under visible light irradiation | Malachite green (MG) | [Dye]=0.4 mm, [Fe3+]: 160 mm, [H2O2]: 440 mm, pH <3 | XTOC: 100% (160 min) | Wu et al. (1999) |
| Photo-Fenton | Azo dye OII | [Dye]=60 mg/l, [H2O2]=5.9 mm, [Fe2+]=0.007 mm, pH 3.0 | XCR: 100%, XDye Removal: 50.5% | Maezono et al. (2011) |
| O3-based AOPs | ||||
| Sequence of ozonation with UV/H2O2 | C.I. direct blue 199 | [Dye]=20.0 mg/l, oxygen flow rate=6.0 L/min, power intensity=560 W | XTOC: 90%, XCR: 100% (95 min) | Shu (2006) |
| Ozonation | Reactive dye and synthetic dye effluent | Stirring speed=500 rpm, average O3 input rate=52.5 mg/l | Deg. rate: 0.03 g/l min (60 min) | Tabrizi et al. (2011) |
| Ozonation | C.I. reactive blue 19 | [Dye]=800 mg/l, T=25°C, O3 dose=55 g/m3 | XCOD: 55%, XTOC: 17% (90 min) | Tehrani-Bagha et al. (2010) |
| Ozonation and biological treatment | Remazol black B dye | [Dye]=500 mg/l, power intensity=50 W, pressure=0.5 kg f/cm2, pH 9, flow rate=1 L/min | XDye Removal: 91% (105 min) | de Souza et al. (2010) |
| Electrochemical oxidation and ozonation processes | Congo red dye | [Dye]=0.5 g/l, current density=30 mA/cm2, O3=1 g/h, V=2 L, pH 12, T=25°C | XCOD: 85%, XTOC: 81% | Faouzi Elahmadi et al. (2009) |
| UV/O3 | Azo dyes | [Dye]=20.0 mg/l, oxygen flow rate=6.0 L/min, O3=110 W and 105 V | XCR: 95% (110 min) | Shu and Chang (2005b) |
| Ozonation | C.I. remazol black 5 | [Dye]=2 g/l, O3 concentration=20.5 mg/l | XCOD: 40%, XTOC: 25% (6 h) | Wang et al. (2003) |
| Ozonation | C.I. reactive orange 122 | [Dye]=0.6 mg/l, [D]=300 mg/l; νEOP=0.25 g/h | XTOC: ≥70% (1 h) | Santana et al. (2009) |
| Electrochemical oxidation | ||||
| Fe2+/Cu2+ combination for the solar PEF treatment | Disperse blue 3 | [Dye]=200 mg, [Fe2+]=0.5 mm, [Cu2+]=0.1 mm, pH 3.0, current density=50 mA/cm2, T=35°C | XTOC: >95% (360 min) | Salazar et al. (2012) |
| EF/BDD | Azo dye | [Dye]=80 mg/l, [Fe2+]=0.3 mm, current density=15 mA/cm2 | XCR: 95.9% (50 min) | Cruz-González et al. (2012) |
| Solar PEF | Disperse red 1 (DR1) and disperse yellow 3 (DY3) | [DR1]=0.520 mm, [DY3]=0.558 mm, [Fe2+]: 0.5 mm, pH 2.0–3.0, current density=50 mA/cm2 | XTOC: >95% (180 min) | Salazar et al. (2011) |
| EF | Textile wastewater | COD=980–1100 mg/l, Fe2+/H2O2=1:10, pH 7.0 | XTOC: 90%, XCR: 86–94% (90 min) | Yu et al. (2013) |
| EF and solar PEF | Acid yellow 36 | [Dye]=108 mg/l, [Fe2+]=0.5 mm, pH 3.0, current=3 A | XTOC: 95% (90 min) | Ruiz et al. (2011a) |
| Solar PEF | Acid red and acid yellow | [Dye]=119 mg/l, [Fe2+]=0.5 mm, pH 3.0, j=50 mA/cm2, Na2SO4=0.05 M | XTOC: 95% (360 min) | Ruiz et al. (2011b) |
| US-assisted EF | RB5 | pH 3, current=0.25 A, H2O2=23.5 mm, electrode distance=2.5 cm | k: 0.9513 | Şahinkaya (2013) |
| Trickle-bed reactor of EF | Reactive brilliant red X-3B (X-3B) | [Dye]=123 mg/l, H2O2=9.43 mm, current=0.3 A | XCR: 97%, XTOC: 87% (3 h) | Lei et al. (2013) |
Utilization of heterogeneous system for degradation of dye via AOPs.
| Method | Catalyst | Pollutant | Operating conditions | Performance | References |
|---|---|---|---|---|---|
| Fenton-based AOPs | |||||
| Photo-Fenton | Pillared laponite clay-based Fe nanocomposites | Acid black 1 | [Dye]=0.1 mm, [H2O2]=6.4 mm, UVC=8 W, pH 3.0 | XColor: 100% (120 min) | Sze Nga Sum et al. (2004) |
| Laponite clay-based Fe nanocomposite (Fe-Lap-RD) | Reactive red HE-3B | [Dye]=100 mg/l, [H2O2]=500 mg/l, catalyst loading=1.0 g/l, pH 3.0, UVC=2×8 W | XTOC: 76%, XColor: 100% (120 min) | Feng et al. (2003) | |
| Fe-laponite | Azo dye acid black 1 | [Dye]=0.1 mm, [H2O2]=6.4 mm, catalyst loading=1 g/l, H2O2/AB1: 64 (molar ratio), UVC=8 W (254 nm), pH 3.0 | XColor: 100%, XTOC: 90–100% (90/120 min) | Sum et al. (2005) | |
| Hydroxyl Fe-pillared bentonite | OII | [Dye]=0.2 mm, [H2O2]=10 mm, catalyst loading=1.0 g/l, pH 9.5 | XTOC: 60%, XColor: 100% (120 min) | Chen and Zhu (2006) | |
| Fe-pillared bentonite (Fe-B) | Azo dye X-3B | [Dye]=0.1 mm, [H2O2]=10 mM, catalyst loading=0.5 g/l, Hg lamp=125 W, pH 3 | XColor: 100% | Li et al. (2006) | |
| Al-Fe-pillared bentonite (Al/Fe-B) | Azo dye X-3B | [Dye]=0.1 mm, [H2O2]=10 mm, catalyst loading=0.5 g/l, Hg lamp=125 W, pH 3 | XColor: 100% | Li et al. (2006) | |
| Bentonite clay-based Fe nanocomposite film | OII | [Dye]=0.2 mm, [H2O2]=10 mm, UVC=1×8 W, pH 3.0 | XColor: 100%, XTOC: 50–60% (120 min) | Feng et al. (2005) | |
| Fe-montmorillonite | MB | [Dye]=0.2 mm, [H2O2]=10 mm, catalyst loading=5 g, pH 3.0 | XDye Removal: 93% (180 min) | De León et al. (2008) | |
| Fe-pillared montmorillonite | Reactive brilliant orange X-GN | [H2O2]=4.9 mm, catalyst loading=0.6 g/l, T=30°C, pH 3.0 | XColor: 98.6%, XTOC: 52.9% (140 min) | Chen et al. (2009a) | |
| Fe-ZSM-5 zeolite | C.I. acid blue 74 | [Dye]=85.6 mm, [H2O2]=21.4 mm, catalyst loading=0.5 g/l, pH 5, UV-C light | XTOC: 57% (120 min) | Kasiri et al. (2008) | |
| Fe/ZSM-5 zeolite | OII | [Dye]=0.1 mm, [H2O2]=6 mm, catalyst loading=200 mg/l, T=53°C, pH 3.0 | XTOC: 90% | Duarte and Madeira (2010) | |
| Fe/C and Fe/Nafion/C catalysts | OII | [Dye]=0.2 mM, [H2O2]=10 mm, pH >4.4, solar irradiation | XTOC: <90% | Gumy et al. (2005) | |
| Fe-Ce bimetal catalyst | Reactive brilliant red X-3B | [Dye]=100 mg/l, [H2O2]=34 mg/l, catalyst loading=0.500 g/l, UV=36 W, pH 3.0 | XColor: <99% (30 min) | Zhang et al. (2007b) | |
| Nafion cation-transfer membranes exchanged with Fe ions | OII | [Dye]=50 mm, [H2O2]=2.42 mm, Suntest=50 mW/cm2, pH 2.8 | XTOC: <70% (30 min) | Fernandez et al. (1999) | |
| Fe-pillared vermiculite | Reactive brilliant orange X-GN | [Dye]=100 mg/l, [H2O2]=3.92 mm, catalyst loading=0.5 g/l, pH 3, T=30°C | XColor: 98.7%, XTOC: 54.4%, RT: 75 min | Chen et al. (2010a,b) | |
| Clay-based Fe nanocomposites | OII | [Dye]=0.2 mm, [H2O2]=10 mm, catalyst loading=1.0 g/l, UVC=1×8 W, pH 3.0 | XColor: 100% XTOC: 90% (60/120 min) | Feng et al. (2006) | |
| Fe-modified local clay | Acid green 25 | [Dye]=50 ppm, [H2O2]=6.7 mm, catalyst loading=1.25 g/l, pH 3 | XColor: 99% (20 min) | Azmi et al. (2014) | |
| Acid-activated Cu-bentonite | Azo dye acid black 1 | [Dye]=0.1 mm, [H2O2]=6.4 mm, H2O2/Dye=64 (molar ratio), catalyst loading=0.5 g/l, UVC light=8 W, pH 3, T=30°C | XDye Removal: 100%, XTOC: 85% (30/120 min) | Yip et al. (2005) | |
| Prussian blue (Fe hexacyanoferrate) colloids | Rhodamine B | [Dye]=0.1 mm, [H2O2]=6.4 mm, H2O2/Dye=64 (molar ratio), catalyst loading=0.5 g/l, UVC light=8 W, pH 3, T=30°C | |||
| Fe-laponite | Reactive red HE-3B | [Dye]=100 mg/l, [H2O2]=14.7 mm, H2O2/Dye=205.3 (molar ratio), catalyst loading: 1 g/l, 2×8 W UVC (254 nm), pH 3.0 | XColor: 100%, XTOC: 76% (30/120 min) | Feng et al. (2003) | |
| Fenton and Fenton-like | ACFs supported 8-hydroxyquinoline ferric | Synthetic dyes | [Dye]=0.01 mM, [H2O2]=30.0 mm, catalyst loading=3.3 g min/ml, pH 2.5 and T=50°C | XDye Removal: >96.6% (10 min) | Sun et al. (2014) |
| Fe/AC (continuous packed-bed reactor) | Alcian blue-tetrakis-AB (methylpyridinium) | [Dye]=0.01 mm, [H2O2]=30.0 mm, catalyst loading=3.3 g min/ml, pH 2.5 and T=50°C | XColor: 93.2%, XTOC: 54.1% | Duarte et al. (2013a) | |
| Fe/AC (continuous packed-bed reactor) | Real textile effluent (cotton industry) | [Dye]=0.01 mm, [H2O2]=30.0 mm, catalyst loading=3.3 g min/ml, pH 2.5 and T=50°C | XColor: 96.7%, XTOC: 73.6%, XCOD: 66.3%, XBOD: 72.5% | Duarte et al. (2013a) | |
| ACFs supported ferric citrate (FeCit@ACFs) | Reactive brilliant red X-3B (X-3B) | [Dye]=50 mm, [H2O2]=20 mm, catalyst loading=5 g/l, pH 7 | XDye Removal: 98.9% (27 min) | Wang et al. (2014b) | |
| AC: Norit RX 3 extra, Merck, and Kynol | Azo dye OII | [Dye]=0.1 mm, [H2O2]=6 mm, catalyst loading=0.1 g/l, pH 3, T=30°C * Fe-Norit | XTOC: 61% | Duarte et al. (2011) | |
| NdFeB magnetic AC-Fenton | Azo dye MO | [Dye]=20 mg/l, T=20°C, catalyst loading=10 g/l, [H2O2]=0.6 mm, pH 3.0 | XTOC: 97.1% | Yang et al. (2014) | |
| Fe2O3-SiO2 | MO | [Dye]=0.6 mg/ml, [H2O2]=2 mL, pH 2.93 | XColor: ≥90% (20 min) | Panda et al. (2011) | |
| AC Norit RX 3 Extra was impregnated with ferrous sulfate | CSB | [Dye]=0.012 mm, [H2O2]=2.25 mm, catalyst loading=4.1 g min m/l, T=50°C, pH 3 | XDye Removal: 88%, XTOC: 47% | Mesquita et al. (2012) | |
| Fe-containing Y and ZSM-5 zeolites | Reactive brilliant blue KN-R | [Dye]=250 mg/l, [H2O2]=30.0 mm, catalyst loading=4.0 g/l, pH 2.5 | XColor: 90% (20 min) | Chen et al. (2008) | |
| US/α-FeOOH/H2O2 | C.I. acid orange 7 | [Dye]=79.5 mg/l, [FeOOH]=0.3 g/l, [H2O2]=7.77 mm, p=80 W/l, pH 3 | XColor: 90%, XTOC: 42% (10/90 min) | Zhang et al. (2009b) | |
| NZVI Fe-Fenton | Azo dye | [H2O2]=33.1–36.8 mm, catalyst loading=225–250 mg/l, pH 3.0 | XCOD: 80% | Yu et al. (2014) | |
| Zeolite Y-Fe catalyst | Azo dye OII | [Dye]=0.1 mm, [H2O2]=6 mm, catalyst loading=200 mg/l, T=30°C, pH 3 | K: 5.83 h-1 | Rache et al. (2014) | |
| Fe-zeolite Y-type catalyst | Acid red 1 | [Dye]=50 mg/l, [H2O2]=16 mm, catalyst dosage=2.50 g/l, pH 2.5, T=30°C | XColor: 99% | Hassan and Hameed (2011b) | |
| NH4ZSM5 and HY zeolite | C.I. reactive blue 49 | Fe2+/H2O2=1: 20, Fe3+/H2O2=1: 10, Fe0/H2O2: 1: 20, c(Fe2+)=0.5 mm, c(Fe3+)=0.5 mm, c(Fe0)=0.5 mm, pH 3, * Fe0/H2O2/HY system best efficiency | XColor: <98%, XTOC: 85% | Kušić et al. (2007b) | |
| NH4ZSM5 and HY zeolite | C.I. reactive blue 137 | Fe2+/H2O2=1: 30, Fe3+/H2O2=1: 40, Fe0/H2O2=1: 20, c(Fe2+)=0.5 mm, c(Fe3+)=0.5 mm, c(Fe0)=1 mm, pH 3 | XColor: <98%, XTOC: 62.6% | Kušić et al. (2007b) | |
| Fe(II) supported on Y zeolite | C.I. acid red 14 | [Dye]=50 ppm, [H2O2]=8.7 mm, [Catalyst]=15 g/l, T=80°C, pH 5.96 | XColor: 99.3±0.2%, XTOC: 84±5% | Idel-aouad et al. (2011) | |
| CuFeZSM-5 zeolite | Rhodamine 6G | H2O2=40 mm, catalyst loading=0.15 g, pH 3.4, T=323K | XColor: 100%, XTOC: 68.8% (45 min/2 h) | Dükkancı et al. (2010) | |
| Fe-doped zeolite Y catalyst/US | Acid red B (ARB) | [Dye]=80 μM, frequency=20 kHz, power=50 W, catalyst dosage=2.0 g/l, pH 3, | XDye Removal: 66% (60 min) | Jamalluddin and Abdullah (2014) | |
| 4A-zeolite-supported α-Fe2O3/US | OII | [Dye]=20 mg/l, catalyst loading=0.5 g/l, pH 6.8 | XDye Removal: 92.5% (80 min) | Chen et al. (2010a,b) | |
| Mesoporous silica Fe-doped catalyst/US | C.I. acid orange 7 | [Dye]=100 mg/l, [H2O2]=8 mm, catalyst loading=0.3 g/l, power=80 W, pH 2 | XColor: 84.9% | Zhong et al. (2011) | |
| Fe-clay | Reactive blue 4 | [Dye]: 50 mg/l, [H2O2]=8 mm, catalyst loading=5.0 g/l, T=30°C, pH 3.0 | XColor: 99% (140 min) | Hassan and Hameed (2011a) | |
| Fe2(MoO4)3 | Acid OII | [Dye]=100 mg/l, [H2O2]=18 mm, catalyst loading=1.4 g/l, pH 6.7 | XTOC: 94.1% (60 min) | Tian et al. (2011) | |
| Fe2MnO4/AC-H, Fe3O4/AC-H | MO | [Dye]=50 mg/l, [H2O2]=18 mm, catalyst loading: 2.5 g/l, T=302.0±1.0 K, pH 4.0±0.1 | XTOC: 59%, XDye Removal: 100% (120 min) | Nguyen et al. (2011) | |
| ZrO2-ZnO/HY nanocomposite | MG, congo red, and MO | [Dye]=10 mg/l, catalyst loading=0.60 g/l, pH 11, visible radiation | XColor: 99.7% (60 min) | Sapawe et al. (2013) | |
| NH4ZSM-5 | C.I. reactive red 45 | [Dye]=80 mg/l, catalyst loading=1 g/l, flow rate=0.1 L/min, T=25±0.2°C | XTOC: 85.4% | Peternel et al. (2007) | |
| Ni-TiO2-modified zeolite | Reactive yellow 125 azo dye | [Dye]=25 mg/l, catalyst loading=1 g/l, pH 3 | XTOC: 91.64%, XColor: 91.60% (240 min) | Ilinoiu et al. (2013) | |
| NiS-P zeolite | Eriochrome black T | [Dye]=40 ppm, catalyst loading=0.8 g/l T=50°C, pH 9.1 | XDye Removal: 91.64% | Nezamzadeh-Ejhieh and Khorsandi (2010a,b) | |
| Ni2+-exchanged zeolite P | Eriochrome black T | [Dye]=30 ppm, catalyst loading=0.8 g/l T=50°C, pH 9 | XDye Removal: <65% | Nezamzadeh-Ejhieh and Khorsandi (2010a,b) | |
| TiO2/clay | Anionic reactive blue 19 | [Dye]=75 mg/l, [H2O2]=7.35 mm, catalyst loading=100 mg/l pH 5.51 | XCOD: 98%, XColor: 100% | Hadjltaief et al. (2014) | |
| Immobilized TiO2 photocatalysis coupling anodic oxidation on BDD electrode | X-3B | [Dye]=50 mg/l, [H2O2]=7.35 mm, current applied=25.2 mA/cm2, Na2SO4=50 mm pH 2.72 | XColor: 100%, XTOC: 74.2% | Zhang et al. (2009a) | |
| ZnO | Acid red 14 | [Dye]=20 ppm, pH neutral, catalyst loading=160 ppm | XColor: 100% | Daneshvar et al. (2004) | |
| TiO2/ZnO nanofibers | Rhodamine B | [Dye]=0.001 mm, catalyst loading=2 g/l, UV lamps=8×8 W | XColor: 100% (100 min) | Pei and Leung (2013) | |
| TiO2/UV | Azo dye acid red 14 | [Dye]=20 ppm, TiO2=40 ppm, pH neutral | K: 1.41×10-2 min-1 (3.5 h) | Daneshvar et al. (2003) | |
| ZnO | Acid brown 14 | [Dye]=0.5 mm, catalyst loading=2.5 g/l, T=25°C and 30°C, pH 3, intensity of sunlight=1.32–1.37×105 | K: 7.48×104 s-1 (360 min) | Sakthivel et al. (2003) | |
| TiO2-montmorillonite | Crystal violet | [Dye]=0.1 mm, catalyst loading: 0.08 g, pH 6–7.5 | XDye Removal: 97.1% | Djellabi et al. (2014) | |
| ZnO photocatalyst supported with porous AC | Alizarin cyanin green | [Dye]=30 ppm, catalyst loading=200 mg, dose of the CAC=40 mg, pH 2 | kt: 0.0465 min-1 (90 min) | Muthirulan et al. (2013) | |
| O3-based AOPs | TiO2-pillared montmorillonite | Solophenyl red 3BL | [Dye]=100 mg/l, pH 5.8 and 2.5 g/l | kt: 0.0096 min-1 | Damardji et al. (2009b) |
| TiO2-pillared Romanian clay | Congo red | [Dye]=20 mg/l, catalyst loading=0.4 g/l, T=25°C, pH 10, RT=30 min | XColor: 100%, XDye Removal: 50% (120 min) | Dvininov et al. (2009) | |
| TiO2 suspension | Procion yellow H-EXL | [Dye]=20 mg/l, catalyst loading=1 g/l, pH 5 | XDye Removal: 100% (120 min) | Barakat (2011) | |
| TiO2 nanoparticles | C.I. basic blue 3 | [Dye]=10 mg/l, UV light intensity=47.2 W/m2, flow rate=100 ml/min | XTOC: 75% (12 H) | Khataee et al. (2010) | |
| γ-Alumina | Remazol brilliant blue R (RBBR) | [Dye]=200 mg/l, T=25°C, stirrer rate=300 rpm; QG=150 L/h, PRO3=2.21 g/h, pH 2.5 | XCOD: 65.6% (30 min) | Erol and Özbelge (2008) | |
| MoO3 nanoparticles | OII | [Dye]=0.1 mm, catalyst loading=250 mg per 250 ml, T=25°C, pH 7.0 | k: 1.656×10-3 S-1 | Manivel et al. (2015) | |
| AC and ceria catalysts | C.I. acid blue 113, reactive yellow 3, and reactive blue 5 | [Dye]=300 mg/l, pH 5.6, catalyst loading=0.5 g/l | XCOD: 63% (120 min) | Faria et al. (2009) | |
| Lanthanum-based perovskites | C.I. reactive blue 5 | [Dye]=1 mm, catalyst loading=100 mg, pH 5.5 | XTOC: 70% (3 h) | Orge et al. (2013) | |
| A porous copper fiber sintered sheet (PCFSS) | Basic yellow 87 | [Dye]=216 ppm, [O3]=500 ppm, T=22°C; pH 6.6, oxygen flow rate=0.5 L/min | XDye Removal: 99%, XTOC: 30%, XCOD: 60% (4 h) | Zhu et al. (2014) | |
| Fe-Cu-O catalyst | ARB | [Dye]=200 mg/l, catalyst loading=1 g/l, pH 6.8 O3 flow=30 mg/min | XCR: 90%, XCOD: 70% (60 min) | Liu et al. (2013) | |
| TiO2/UV/O3 | Brilliant red X-3B | [Dye]=0.2 mm, TiO2=20 mg, O3 flow=0.8 L/min | XDye Removal: 91% (180 min) | Sun et al. (2013) | |
| Electrochemical oxidation-based AOPs | MgO nanocrystals | Reactive red 198 | [Dye]=200 mg/l, dose of O3=0.2 g/h, pH 8, MgO dosage=5 g/l | XDye Removal: 61%, XCR: 99% (10 min) | Moussavi and Mahmoudi (2009) |
| Carboxylated CNTs | Indigo | [Dye]=100 mg/l, [CNTs]=8 mg/l, [AC]=8 mg/l, pH 4.0 | XTOC: 31.9% (2 h) | Qu et al. (2015) | |
| UV/TiO2/O3 | RR2 | [Dye]=100 mg/l, TiO2=1 g/l, pH 6.9 | k: 0.815 min-1 | Wu et al. (2008c) | |
| TiO2-coated stainless steel electrode | Azo dye orange G | [Dye]=64.0 mg/l, Na2SO4=0.01 M pH 3.0, power density=330 μW/cm2 | XDOC: 21.2% | Lin et al. (2013) | |
| Mesoporous silica SBA-15-supported Fe and Co catalysts (Fe-Co/SBA-15) | OII | [Dye]=100 mg/l, PDS=2.0 g/l, catalyst=1.0 g/l, j=8.40 mA/cm2, pH 6, Na2SO4=0.05 M, T=20°C | XCOD: 52.1%, XTOC: 31.9% (60 min) | Cai et al. (2014) | |
| Fe2O3/γ-Al2O3 | Acid red 3R | [Dye]=500 mg/l, pH 3.0, airflow=0.20 m3/h voltage=20.0 V, interelectrode distance=3.0 cm | XCR: 77.2% (100 min) | Yue et al. (2014) | |
| TiO2 nanoparticles | C.I. basic red 46 (BR46) | [Dye]=15 mg/l, [Fe3+]=0.1 mm, applied current=300 mA | XTOC: 99.2% (35 min) | Khataee et al. (2011b) | |
| Immobilized ZnO nanoparticles | C.I. direct yellow 12 (DY12) | [Dye]=50 mg/l, Na2SO4=0.05 M applied current=100 mA, pH 3 | XTOC: 92.7% (90 min) | Khataee and Zarei (2011a) | |
| TiO2 nanoparticles and CNT cathode | C.I. acid green 25 | [Dye]=10 mg/l, UV light intensity=47.2 W/m2, flow rate=100 ml/min | XCR: 78.12%, XTOC: 85.5% (6 h) | Khataee et al. (2011a) | |
| ZnO nanophotocatalyst and CNT-based cathode | DY12 | [Dye]=50 mg/l, [Fe3+]=0.2 mm, applied current=400 mA, Na2SO4=0.05 M, pH 3.0 | XTOC: 96.7% (6 h) | Khataee and Zarei (2011b) | |
| TiO2 nanoparticles | BR46 | [Dye]=0.1 mm, catalyst loading=15 mg/l, applied current=300 Ma | XCR: 88.9%, XTOC: 99.2% (35 min) | Khataee et al. (2011b) | |
| TiO2 nanoparticles and CNT cathode | BR46 | [Dye]=20 mg/l, I=100 mA, pH 3.0, Na2SO4=0.05 mol/l, [Fe3+]0=0.1 mmol/l | XTOC: 98.8% (6 h) | Zarei et al. (2010) | |
| Photocatalytic oxidation | Nickel-dimethylglyoxime/ZSM-5 zeolite | Methyl green | [Dye]=40 ppm, pH 9, T=60°C, catalyst=0.6 g/l | XCR: >90% | Nezamzadeh-Ejhieh and Shams-Ghahfarokhi (2013) |
| ZnO onto clinoptilolite nanoparticles (NCP) | Bromothymol blue | [Dye]=4 mg/l, pH 7.8, catalyst=0.1 g/l | k: 4.75×10-3 min-1 (240 min) | Bahrami and Nezamzadeh-Ejhieh (2015) | |
| Fe(II)-o-phenanthroline/zeolite Y nanocluster | Methyl green | [Dye]=40 ppm, catalyst=1 g/l, pH 9 | XUV: 80% (120 min) | Nezamzadeh-Ejhieh and Shahriari (2011) | |
| Ni/P zeolite and NiS/P zeolite | Azo dye | [Dye]=20 ppm, catalyst=0.8 g/l, pH 5 | XCR: >85% | Nezamzadeh-Ejhieh and Khorsandi (2011) | |
| ZrO2 nanoparticles | Victoria blue (VB) and direct red (DR) | [Dye]=10 mm, catalyst=30 mg, pH 8 (VB) and 4 (DR) | k: 0.12867 min-1 | Bansal et al. (2015) | |
| Nano-strontium titanate | Azo dyes | [Dye]=20 mg/l, T=25°C, perovskite=0.1%, UV irradiation=20 and 400 W | XCR: >90% | Karimi et al. (2014) | |
| Microsized TiO2 as coated on porcelain-grès tiles | Rhodamine B (RhB), MB, and crystal violet (CV) | [Dye]=1 mm, TiO2=0.1 g/l, UV-A lamp=125 W | XCV: 100%, XMB: 100%, XRhB: 71% | Bianchi et al. (2014) | |
| CuO nanostructures | MB and methylene violet (MV) | [Dye]=15 mg/l, CuO=5 mg | XMB: 89%, XMV: 96% (30 min) | Sonia et al. (2015) | |
DOC, Dissolved organic compound.
4.1 Operating pH
Each AOP only works efficiently and economically within a range of pH. For example, ozonation requires higher pH (alkaline) for unselective reaction with all the recalcitrant contaminants (Ikehata and El-Din 2004, Vogna et al. 2004, Chen et al. 2009b, Beltrán et al. 2012, Loeb et al. 2012). In contrast, the catalytic decomposition of H2O2, by means of Fe salts (Fenton-like process), requires very low pH, giving rise to ·OH (Alaton and Teksoy 2007, Babuponnusami and Muthukumar 2013). The optimum pH for Fenton reaction is approximately 3 regardless of the target substrate (Kusic et al. 2006, Bagal and Gogate 2014). At higher pH, Fe3+ form Fe(OH)3, which reacts slowly with H2O2 (Babuponnusami and Muthukumar 2013). This process may decrease the efficiency of a Fenton system, as less Fe3+ is present to react with H2O2 to generate ·OH. Besides, the autodecomposition of H2O2 also accelerates at higher pH. At very low pH, Fe complex, [Fe(H2O)6]2+, is present and it reacts with H2O2 in the solution. This reduces the amount of Fe2+ present in the solution (Navarro et al. 2010). In addition, at very low pH, H2O2 forms stable oxonium ions [H3O2]+, which are stable and less reactive compared to ·OH, reducing its efficiency in oxidizing the pollutants.
Besides, pH is one of the most important operating parameters that control the photocatalytic activity. This is because pH has strong effects on the catalyst’s surface charge, flat band potential, and dissociation of compounds in the solution (Nezamzadeh-Ejhieh and Amiri 2013). These multiple roles have made the interpretation of pH effects on the efficiency of photocatalytic degradation difficult. The point of zero charge (pHPZC) is the pH at which the surface of a catalyst is uncharged. Catalyst surface is positively charged at pH values smaller than pHPZC and negative at higher values (Nezamzadeh-Ejhieh and Zabihi-Mobarakeh 2014). Most of the study investigated the amplitude pH in the range of 2–12 (Nezamzadeh-Ejhieh and Khorsandi 2014). It is activity largely dependent on the acid base property of the catalyst’s surface and can be described based on the PZC (Adam et al. 2011). At very acidic condition (pH 2–3), the acidification of the solution using HCl generates much HClO·- from the reaction between Cl- and ·OH. As a result, the degradation efficiency is decreased due to the lower reactivity of HClO·- compared to ·OH (Nezamzadeh-Ejhieh and Khorsandi 2010a,b, Pourtaheri and Nezamzadeh-Ejhieh 2015). It is therefore rational to assume that the pH of a working solution is based on the charge of the catalyst (Muruganandham and Swaminathan 2006b).
Muruganandham and Swaminathan (2006a) studied the effect of pH on the degradation and decolorization efficiency of Reactive Orange 4. Their study reported that the adsorption of dye on TiO2 surface is reduced with increased pH. The catalyst used in their study has a PZC of 6.8. In acidic solution, the pH is lower than the PZC; hence, TiO2 is positively charged. On the contrary, in alkaline solution, TiO2 is negatively charged. In their study, dye was adsorbed onto TiO2 when the pH was near PZC (6.8) due to the strong adsorption between the negatively charged dye and the positively charged TiO2 surface. The negatively charged dye was repelled at pH above 6.8, as the surface of TiO2 was negatively charged, reducing the adsorption efficiency. A similar result was obtained by Balcioglu and Arslan (1997), Muruganandham and Swaminathan (2006b), and Park et al. (2003) for the degradation of dyes. In addition, some investigations have shown that photoreactivity increases in alkaline medium. The highest reaction rate was achieved for the degradation of monoazo dye Procion Red MX-5B at pH 10 for 100 mg/l TiO2 in a study conducted by So et al. (2002). On the contrary, Arabpour and Nezamzadeh-Ejhieh (2015) showed different outcomes in comparison with other researchers, as best degradation was achieved at very acidic condition (pH 3) for the catalyst with pHPZC approximately 5.8. This is might be due to the fact that the strong adsorption of pollutant on the catalyst plays a positive role to enhance the degradation rate at stronger acidic conditions. In a study conducted by Nezamzadeh-Ejhieh and Zabihi-Mobarakeh (2014), maximum decolorization was observed at pH 5.9, and decreases in the photodecolorization extent were observed at pH 9.5. This is might be due to the repulsive forces between the negatively charged surface and anionic dye molecules. As a conclusion, it clearly shows that the pH of photocatalytic process depends on the catalyst applied, pollutants, and oxidants used.
4.2 Initial concentration of oxidant
The degradation rate of dye wastewater is also affected by the initial concentration of oxidants. H2O2 is a very important source of ·OH. Previous studies have shown that degradation efficiency increases with H2O2 concentration up to a threshold value, beyond which degradation efficiency significantly declines. This is because H2O2 scavenges ·OH at higher concentration, and this is applicable for any systems that involve H2O2 as an oxidant, such as TiO2/UV/H2O2, UV/H2O2, O3/H2O2, Fenton, and photo-Fenton (Schulte et al. 1995, Kusic et al. 2006, Yu et al. 2010, Torrades and García-Montaño 2014). At higher concentration, H2O2 reacts with ·OH and generates hydroperoxyl radicals, which are less reactive compared to ·OH (Meric et al. 2005, Nidheesh and Gandhimathi 2012). In addition, the recombination of ·OH also reduces the degradation efficiency, as fewer radicals are present for oxidation (Neyens and Baeyens 2003). H2O2 can cause an increase in COD and is costly at the same time. It is therefore important to optimize the concentration of H2O2 to avoid its harmful effects and reduce the cost of the system. Besides H2O2, electron acceptors, such as KBrO3, are added in photocatalytic oxidation to enhance the generation of ·OH, increase the rate of reaction, and avoid the electron-hole recombination (Nezamzadeh-Ejhieh and Moazzeni 2013). Previous studies have shown that the rate of pollutant decomposition increases with the concentration of KBrO3 (Nezamzadeh-Ejhieh and Moeinirad 2011). KBrO3 is reduced by the mechanisms shown in Equations (29) and (30) and can decrease electron-hole recombination, which is the major problem in the photocatalytic processes. Bromate ions act as an oxidizing agent and produce other oxidizing agents such as BrO2- and HOBr, although it prevents the direct production of ·OH as shown in Equations (21) and (22) (Amiri and Nezamzadeh-Ejhieh 2015).
4.3 Initial catalyst concentration
Initial catalyst concentration is another factor that affects wastewater treatment efficiency. A number of studies have shown the dependence of photocatalytic degradation rate on the initial concentration of catalyst. Photocatalytic efficiency is found to increase with catalyst loading up to certain dosage and further increase results in a reduction of the removal rate (Wu 2008, Bouanimba et al. 2011, Nezamzadeh-Ejhieh and Karimi-Shamsabadi 2014). At a certain point, although catalyst loading increases the number of active site, all the beam energy is attenuated within the reactor. Further increase in the catalyst loading results in nonuniform light intensity distribution and causes decreases in the reaction rate (Muruganandham and Swaminathan 2006b, Chen et al. 2013). For photocatalytic systems, an optimum catalyst concentration of approximately 1 g/l has been reported in many studies. In addition, a similar scenario can also be observed in Fenton-based processes and homogenous AOPs. Increased degradation rate is usually observed with increased concentration of Fe2+ to a certain level, above which degradation becomes insignificant. For homogenous systems, an excessive amount of unused Fe salts leads to sludge production and post-treatment (Kušić et al. 2007a, Arslan-Alaton et al. 2009, Hammami et al. 2012). Therefore, it is important to optimize the concentration of Fe salts for each treatment system to minimize the toxicity and post-treatment cost.
4.4 Initial concentration of pollutants
Many studies have reported that the initial concentration of pollutants has a significant effect on the degradation efficiency of AOP treatment systems (Tang and Huren 1995, Mohammadi and Sabbaghi 2014). In photocatalytic systems, increasing the organic pollutant concentration significantly enhances the collision between organic contaminants and ·OH, which increases degradation efficiency. However, increasing the concentration more than the optimal point showed decreases in the degradation efficiency (Nezamzadeh-Ejhieh and Moeinirad 2011). This might be due to the fact that more organic compounds are absorbed on the surface of photocatalysts and this reduces the active sites of the catalysts for photons (El-Bahy et al. 2008, Nezamzadeh-Ejhieh and Shahriari 2014). Besides, an increment in the dye concentration also leads to an increase in the equilibrium adsorption of dyes at the active sites of the catalyst surface. More compounds get absorbed onto the surface of photocatalysts when the targeted pollutant increases. Consequently, it decreases the competitive adsorption of O2 and OH- on the active sites of catalysts and decreases the generation rate of O2·- and ·OH (Bahrami and Nezamzadeh-Ejhieh 2015). As a result, there is an inadequate amount of ·OH species to treat highly concentrated wastewater and thus reduce the degradation efficiency. Besides, dye degradation rate decreases with increasing dye concentration due to interference from the intermediates formed upon the degradation of the parental dye molecules that get trapped on the surface of catalysts and causes the deactivation of active sites. UV-based homogenous and heterogeneous AOPs are also not efficient in treating wastewater with high pollutant concentration due to the insufficient penetration of UV radiation. An increase in the pollutant concentration results in the reduced path length of photons entering the solution and this ultimately reduces the number of photons absorbed on the catalysts and decreases the rate of degradation (Bahrami and Nezamzadeh-Ejhieh 2015). In addition, an increase in dye concentration in ozonation systems reduces the color and total organic carbon (TOC) removal at steady state, which means that O3 consumption potentially increases in this situation (Soares et al. 2006, Mahmoodi 2013).
4.5 Effects of salts
The inorganic anions present in wastewater affects the degradation of pollutants. These ions attach on the active sites of catalysts and decrease dye degradation efficiency (Arslan-Alaton 2003). In addition, the presence of inorganic anions increases the concentrations of active radicals and their capabilities for the efficient oxidation of organic pollutants existing in wastewater (Merouani et al. 2010). The radical anions that are formed through this process also behave as radical scavengers, resulting in prolonged dye degradation. They react with ·OH, resulting in high requirements of ·OH species for degradation. The commonly found inorganic anions in wastewater are NaCl, KNO3, NaNO3, Na3PO4, Na2SO4, Na2CO3, and NaHCO3 (Riga et al. 2007, Devi et al. 2011). Inorganic ions affect the efficiency of AOP systems through the formation of Fe3+ complex, generation of toxic products, ·OH scavenging, reaction with H2O2, competition with organic compounds for active sites, sludge formation, and Fe salt recovery (Riga et al. 2007, Orozco et al. 2008). NaCl is used to enhance dye diffusion and its adsorption on fibers, whereas bicarbonate and carbonate are used for dye fixation through covalent bonds ions (Aleboyeh et al. 2012). They are therefore commonly found in colored solution.
4.6 Intensity of radiation
UV irradiation can significantly influence the direct formation of ·OH and the photoreduction rate of Fe3+ to Fe2+ (Papic et al. 2009). The reaction rate is directly proportional to the intensity of radiation, but the reaction rate decreases beyond a certain magnitude of intensity. The application of radiation in AOPs helps to increase decolorization and mineralization efficiency. The decolorization efficiency of the system also depends on both the reactivity and the photosensitivity of the dye. Because most of the dyes are light resistant, it is necessary that UV radiation is combined with powerful oxidants such as H2O2 to enhance the photolysis process [Equation (23); Dükkancı et al. 2014].
Besides, irradiation wavelength is important in an irradiated system. It should be noted that shorter wavelengths are recommended for better removal efficiency.
5 Advantages and limitations of conventional AOPs
In the past 30 years, the efficiency of AOPs in treating different recalcitrant wastewaters containing refractory and toxic pollutants has been studied (Balcıoğlu et al. 2003). AOPs and various combinations of AOPs are known to generate strong ·OH. These processes increase the efficiency of decomposition of impurities present in textile wastewater and reduce the cost of the process.
Previous researches have proven that AOPs are effective for wastewater treatment. AOPs offer several advantages such as the effective removal of organic compounds and the complete mineralization of organic compounds into H2O and CO2. They have low selectivity, which helps solve a lot of pollution problems. Besides, AOPs can generate ·OH through various ways and can be used as a pretreatment method to fully or partially degrade nonbiodegradable pollutants before conventional biological treatments to reduce the treatment cost. AOPs are more frequently used to degrade contaminants that are resistant to conventional treatment. However, AOPs are ineffective due to the high consumption of reagents, especially H2O2. Besides, homogeneous Fenton-based AOPs generate contaminated intermediate products and sludge, which require secondary treatment. Table 6 summarizes the advantages and limitations of most commonly applied AOPs for wastewater treatment.
Advantages and limitations of conventional AOPs.
| AOP | Advantages | Limitation | References |
|---|---|---|---|
| O3 | Strong oxidation power; easily performed; short reaction time; no sludge remains; all residual of O3 easily decomposed | Higher operation cost; energy-consuming process; needs pretreatment | Ikehata and El-Din (2004), Fanchiang and Tseng (2009), and Abu Amr and Aziz (2012) |
| O3/UV | Higher efficiency; more efficiency at generating ·OH; more effective than O3 alone or UV alone | Not cost-effective; energy intensive; mass transfer limitation; sludge production; turbidity can interfere with the penetration of light | Guittonneau et al. (1990), Ruppert et al. (1994), Popiel et al. (2008), Lucas et al. (2010), and Rao and Chu (2010) |
| H2O2/UV | Disinfectant; simple process | Turbidity can interfere with the penetration of light; less efficient in generating ·OH | Elmorsi et al. (2010), Hu et al. (2011), and Karci et al. (2012) |
| O3/H2O2 | More efficient than O3 or H2O2 alone; effective for H2O treatment | Bromate formation; excess usage of H2O2; not cost-effective; energy-intensive | Latifoglu and Gurol (2003), Ikehata et al. (2006), and Lanao et al. (2008) |
| O3/H2O2/UV | Nonselectively with all species in solution; degradation of aromatics and polyphenols was found to be significantly faster | Expensive; COD removal not complete; sludge production | Monteagudo et al. (2005) and Lucas et al. (2010) |
| Fenton-based process | Rapid reaction rates; small footprint; cost and energy effective; generate strong ·OH; degrade a wide range of recalcitrant components; no mass transfer; recycling of ferrous catalysts by reduction of Fe3+ | Not full-scale application exists; small of sludge production; acidic environment | Zhang et al. (2005), Park et al. (2006), Pignatello et al. (2006), Giroto et al. (2008), and Catrinescu et al. (2012) |
| Electrochemical-based processes | Environment friendly; effective one-electron oxidizing agent ·OH; amenability to automation | Cost- and energy-intensive; electrode reliability | Anotai et al. (2006), Huang et al. (2008), Nidheesh and Gandhimathi (2012), and Su et al. (2012) |
6 Advancement in AOPs
AOPs offer several advantages over conventional biological or physical processes, but higher operating cost and chemical consumption limit their industrial applications. UV radiation is one of the AOPs that are available commercially and widely used for disinfecting groundwater and drinking H2O. Other processes, such as Fenton, ozonation, combination of H2O2, UV, O3, photocatalytic oxidation, and US, have been developed at the laboratory, pilot scale and full scale. Current researches mostly focus on improving AOPs to enhance the performance of conventional AOPs to improve their cost efficiency. The following section discusses the recent advancements in AOPs, which improve their feasibility for wider industrial applications.
6.1 Catalytic oxidation
Catalysts have been found as an important tool for waste minimization and pollution prevention. A catalyst is commonly defined as a substance that changes the rate of chemical reaction without being substantially consumed or changed in the process (Regalbuto 2006). Catalysts accelerate reaction rate toward chemical equilibrium to improve the process economically by decreasing activation energy. Catalysts are also known as “green chemicals” that reduce the use and generation of hazardous substances. A careful selection of catalysts allows us to complete a chemical process with zero waste. The commercial feasibility and inherent greenness of any catalyst depend on the selectivity, turnover frequency, and turnover number of that particular catalyst (Lancaster and Chemistry 2010, Suib 2013). The use of catalysts has attracted much attention from researchers because it shows the biggest potential of advancement. Literatures show that various catalysts can be used in different conventional AOPs to minimize the operating cost. To date, different catalysts have been identified as potentially useful for AOPs in waste management.
Catalysts are generally divided into two types: homogenous and heterogeneous catalysts. In a homogeneous reaction, the catalysts are in the same phase as the reactants and they are uniformly distributed within the reaction medium. Therefore, the reaction takes place within the liquid, as the catalyst is dissolved in the reaction medium. Heterogeneous catalysts are used in a phase different from the reactants and the reaction occurs at gas-solid or liquid-solid interfaces (Spivey et al. 2005). A heterogeneous catalyst is also referred to as a surface catalyst, as the reactions take place on the surface of the catalysts, externally or internally within the pores of the catalysts.
6.1.1 Homogeneous catalytic oxidation
Homogeneous catalysts are well researched, as their catalytic activity is easily understood. Transition metals have been used extensively as a homogenous catalyst in a wide range of industrial processes. Examples of AOPs that successfully use homogenous catalysts for pollution degradation include Fenton, photo-Fenton, and O3-based processes (Beltrán et al. 2005, Alaton and Teksoy 2007). The most widely used transition metal as a homogeneous catalyst in Fenton and O3-based AOPs are Fe2+, Fe3+, Zn2+, Mn2+, Mn3+, Mn4+, Ti2+, Cr3+, Cu2+, Co2+, Ni2+, Cd2+, and Pb2+ (Ksenofontova et al. 2003, Abd El-Raady and Nakajima 2005, Beltrán et al. 2005, Alaton and Teksoy 2007, Bautista et al. 2007, Kušić et al. 2007a, Azbar et al. 2008, Zhu et al. 2011, Leong and Bashah 2012, Torrades and García-Montaño 2014).
Fenton and photo-Fenton are inexpensive and easy to operate compared to other AOPs (Kusic et al. 2006, Arslan-Alaton et al. 2009). Fe salts are commonly used as a source of catalysts to generate ·OH in a classical Fenton-based process. A number of critical reviews based on Fenton oxidation for dye removal are available in the existing literatures (Neyens and Baeyens 2003, Gogate and Pandit 2004a, Pignatello et al. 2006, Babuponnusami and Muthukumar 2013, Bagal and Gogate 2014). Conventional Fenton-based processes are commonly integrated with O3, UV, US, electrochemical oxidation, and Fe free catalyst to enhance the degradation efficiency of recalcitrant organic compounds (Malik and Saha 2003, Gogate and Pandit 2004a, Muruganandham 2004, Torrades et al. 2004, Saravanan and Sivasankar 2014). However, homogenous Fenton-based reaction uses a large amount of Fe salts and generates Fe sludge that needs post-treatment, which limits the applicability of Fenton process at industrial scale.
The predominantly used transition metal for homogeneous catalytic ozonation is Fe2+. In a catalytic O3/Fe2+ system, Fe2+ reacts directly with O3 to produce ·OH [refer to Equations (24)–(26); Sauleda and Brillas 2001]:
The initial degradation efficiency of catalytic ozonation (O3/UV/Fe2+) can be improved using Fe salts, but the degradation process can be inhibited by the formation of Fe3+ complexes with intermediate products that can be photodecomposed as reported by Sauleda and Brillas (2001). Based on the literature, the degradation efficiency of pollutants is improved using Fe salts at acidic pH compared to noncatalyzed ozonation (Hammad Khan and Jung 2008). However, Fe2+ ozonation requires less Fe salts to avoid the scavenging effects caused by Fe3+. This is found to be one of the limitations of homogeneous catalysts.
Besides, several studies have also reported that O3/Mn(II) is effective in degrading organics. Wu et al. (2008b) elucidated the decolorization of C.I. Reactive Red 2 (RR2) by different types of homogeneous catalytic ozonation systems, including Mn(II), Fe(II), Fe(III), Zn(II), Co(II), and Ni(II). The authors concluded that the decolorization reaction in catalytic ozonation systems was mainly of radical-type mechanisms, except in the O3/Mn(II) system. The decolorization efficiency of catalytic ozonation exceeded that of ozonation alone and the decolorization efficiency was higher than TOC removal efficiency in all the experimental conditions.
It should be noted that the generation of Fe sludge, short life span of catalysts, difficulty in separating catalysts from the treated solution, presence of scavenger, narrow range of pH, possible regeneration of catalyst, formation of by-products, high post-treatment cost, and requirement of a large amount of oxidants are the main disadvantages of homogeneous catalysts in AOPs. Therefore, it is necessary to develop various catalytic systems for efficient wastewater treatment. Table 4 summarizes the relevant experimental works and operating parameters involved in various types of homogeneous AOPs. Based on the literatures, it can be concluded that homogenous systems are highly affected by the concentration and type of catalysts, pH of the treatment system, light intensity, type of pollutant, concentration of oxidants, presence of ion species, and type of intermediates.
6.1.2 Heterogeneous catalytic oxidation
Heterogeneous catalysts can overcome the shortcoming of a homogeneous system, as they are capable of inhibiting the generation of Fe sludge. This type of catalyst is known to be greatly effective in treating industrial wastewater containing nonbiodegradable contaminants, as it has high catalytic activity. In a heterogeneous system, soluble Fe2+ is replaced by Fe-containing solids. It involves an intensive contact of an organic compound in a solution with solid catalysts. The advantages of heterogeneous Fenton over Fenton process include high degradation activity under a broad range of pH, low cost, no Fe sludge formation, easy separation of catalyst, high stability, and improved adsorption capacity (Garrido-Ramírez et al. 2010, Tizaoui et al. 2011, Catrinescu et al. 2012). As noted, Fenton process is divided into two stages. At the first step, the generated ·OH decomposes very rapidly and the oxidation rate is slower due to the slow regeneration of Fe2+ to Fe3+. Heterogeneous catalysts are attractive, as they are capable of controlling the formation and consumption of Fe ions during oxidation. Besides, high dispersion and large surface area of some catalysts offer high degradation efficiency. Although the incorporation of catalyst makes oxidation reaction milder than classical Fenton process, it requires moderate temperature and pressure (Dantas et al. 2006).
An ideal heterogeneous catalyst should be applicable in a wide range of pH, cost-effective, and photosensitive and has high catalytic activity but low catalyst leaching. Various FeO and transition metals have been used as heterogeneous catalysts to effectively degrade recalcitrant organic compounds. FeO plays a significant role in controlling the concentration, migration, and conversion of pollutants in the surface environment. Heterogeneous catalysts that have been employed until now in AOPs include photocatalysts (TiO2), ferrihydrite, α-Fe2O3, α-FeOOH, lepidocrocite (γ-FeOOH), magnetite, pyrite, and activated carbon (AC; Matta et al. 2007, Rusevova et al. 2012). The incorporation of transition metals on different supports such as Nafion, metals, AC, graphene, carbon aerogel, carbon nanotube (CNT), clays, polymers, alumina, fly ashes, and zeolite have been proven to have excellent catalytic activity in heterogeneous AOPs for wastewater treatment (Dantas et al. 2006, Gonzalez-Olmos et al. 2012). In heterogeneous reaction, the active phase is supported by porous matrices. The selection of catalyst supports is very important during the preparation of heterogeneous catalysts. Immobilization of catalysts over different supports enables the application of catalysts within a wider pH range. The nature and properties of a support play a crucial role in modulating the activities in the catalytic sites. Reviews on the other heterogeneous catalysts to treat wastewater have been reported by a few researchers. The following section discusses the most applied heterogeneous catalysts for different types of AOPs. Table 5 summarizes the relevant experimental work and operating parameters involved in various types of heterogeneous AOPs. Based on the literatures, it can be concluded that heterogeneous systems are highly affected by types and characteristics of catalysts, pH of the treatment system, light intensity, type of pollutant, concentration of oxidants, presence of ion species, and type of intermediates.
6.1.2.1 FeO minerals
The effectiveness of FeO as a heterogeneous catalyst to degrade a wide range of pollutants is available in the literatures (Huang et al. 2001, Baldrian et al. 2006, Dantas et al. 2006, Hanna et al. 2008, Pham et al. 2012). Heterogeneous FeO have been used as an efficient catalyst with a narrow band gap (2.0–2.3 eV). The most commonly used FeO for pollutant degradation in the literatures are α-FeOOH, α-Fe2O3, ferrihydrite, and magnetite (Hanna et al. 2008). The types and surface areas of FeO, concentration of H2O2 and Fe salts, pH of solutions, and properties of pollutants are the main parameters that affect the mineralization of organic pollutants. Transition metals substitute FeO to provide a new promising class of heterogeneous catalysts that effectively degrade recalcitrant organic compounds in AOP systems. Transition metals are used due to low cost, easy magnetic separation, high stability, and improved adsorption capacity. Besides, the coupling and supporting of semiconductors with different band-gap energy levels causes a significant increase in the photocatalytic efficiency of FeO (Aleksić et al. 2010). The study conducted by Arabpour and Nezamzadeh-Ejhieh (2015) shows that supporting FeO semiconductor onto clinoptilolite nanoparticles significantly increases its photocatalytic reactivity, as more active sites of the catalyst are available for photons and it significantly generates more ·OH. Besides, by supporting FeO with zeolite nanoparticles, it was found to increase the adsorption and ion-exchange capability of zeolites (Nezamzadeh-Ejhieh and Ghanbari-Mobarakeh 2015). It should be noted that hybridizing and supporting FeO with other semiconductors significantly increase the photocatalytic efficiency by decreasing the electron-hole recombination in comparison with the semiconductor alone (Bahrami and Nezamzadeh-Ejhieh 2014, Mousavi-Mortazavi and Nezamzadeh-Ejhieh 2015).
α-FeOOH is a common oxy-hydroxide in the environment and found in association with α-Fe2O3 (Auerbach et al. 2003). α-FeOOH is the most abundant and stable among all and it possesses the capability of incorporating a wide range of environmentally important oxy-anions and cations in its complex matrix. α-FeOOH is commonly used for the degradation of organic pollutants among other FeO, as it is workable in a wide pH range, relatively cheaper than the others, environment friendly, and thermodynamically stable and has a higher efficiency under UV radiation (Li et al. 2007).
In the Jaiswal et al. (2013) study, α-FeOOH was used as an adsorbent for the removal of heavy metals such as Cu2+ and Cd2+. These toxic metals can be actively found in textile wastewater and they are carcinogenic and nonbiodegradable and can contribute to bioaccumulation. Some experimental results have shown that the percentage adsorption of Cu(II) and Cd(II) are 98.00% and 87.50%, respectively, and depicted by pseudo-second-order kinetic equations. In addition, the photodegradation of an azo dye, Mordant Yellow 10, in aqueous dispersions of α-FeOOH at neutral pH under UV irradiation has been investigated (He et al. 2002). The work by He et al. (2002) provided a promising way for the degradation of nonbiodegradable contaminants at neutral pH. Intermediate products, such as acetic acid, nitrobenzene, and 4-hydroxybenzenesulfonic, which are formed during the degradation process, indicate that organic pollutants are brought into contact with the aqueous solutions of H2O2 in the presence of α-FeOOH particles. The study concluded that the α-FeOOH/H2O2 system has more advantages than the homogeneous Fenton reagent Fe2+/H2O2 from the viewpoint of the reaction pH range and the removal of Fe ions.
Besides, the ionic substitution of metals such as Al for Fe in the structure of α-FeOOH improves the solubility, crystallization rate, and surface magnetic properties of α-FeOOH (Wang et al. 2014a). Therefore, many works have been conducted on developing a new catalyst based on α-FeOOH, as it is cost efficient and environment friendly. The S-containing grinding wheel ash has been evaluated for the decolorization of methylene blue (MB) in the presence of H2O2 at various pH conditions (Yuan et al. 2013). In the study by Yuan et al. (2013), Fenton-like process was found to effectively oxidize MB at neutral pH and 72% of TOC removal could be achieved after 60 min of catalytic oxidation. Silica enhances the α-FeOOH adsorptive capacity by inhibiting the formation of the crystalline structure of FeOOH and the surface catalytic oxidation process through a nonradical mechanism. In addition, alumina enhances the α-FeOOH catalytic activity and prevents the Fe from leaching.
Fe3O4 magnetite nanoparticles have also been widely reported as a suitable heterogeneous catalyst among other FeO. This is due to their intrinsic peroxidase-like activity that accelerates the decomposition of H2O2 and easy separation from the reaction medium by an external magnetic field (Chen et al. 2010a,b). Catalytic activities of magnetite can be improved by substituting Fe with other cations (transition metals), composite formation, or immobilization of metals on the magnetite surface. A study was conducted by Liang et al. (2012) to investigate the catalytic activity of chromium-substituted magnetite in the heterogeneous Fenton degradation of cationic and anionic dyes. In their study, the Brunauer-Emmett-Teller (BET) surface area and superficial hydroxyl amount of magnetite was found to increase with substitution of Cr3+, which consequently improved the adsorption ability of dyes. It was noted that the metal-substituted magnetite was efficient in treating textile wastewater and environment friendly at the same time. The authors conducted a further study by investigating the efficiency of Ti4+- and V3+-substituted magnetite in Fenton catalytic activity to treat dye wastewater (Liang et al. 2013). They reported that all the transition metals improved the degradation of organic pollutants by accelerating the OH free radical generation and enhancing the adsorption of pollutant and H2O2. In addition, the efficiency of decolorization of MB by heterogeneous UV-Fenton reactions catalyzed by V-Ti co-doped magnetites has been studied (Liang et al. 2012). The contribution of V and Ti cations on improving the adsorption and catalytic activity of magnetite was also highlighted in the study conducted by Liang et al. (2012). The study concluded that UV greatly accelerated MB degradation and magnetite with more Ti cation and less V cation. Therefore, it was noted that Ti4+ was more prominent in improving the adsorption and catalytic activity of magnetite than V3+. Magnetite is widely used in Fenton reaction due to its magnetic and redox characteristics. However, the main problem associated with the use of magnetite is the production of Fe3+ oxide layer during oxidation, which reduces the catalytic activity of magnetite to produce ·OH for the oxidation process. Therefore, it is important to develop more stable and efficient magnetite catalysts to efficiently oxidize H2O2 to generate reactive free radicals for the oxidation purpose. Magnetite is sensitive to high temperature and can be oxidized to α-Fe2O3 when heated above 600°C by simple heating with air.
The activity of α-Fe2O3 can also be improved by incorporating different metals into the structure, as they can increase the catalytic activity and stability of the catalyst. The photocatalytic activity of α-Fe2O3 increases with its crystallinity. High temperature is needed to obtain highly crystallized α-Fe2O3 that can also render the growth of α-Fe2O3 particles at the same time. The removal efficiency of MB by niobium-incorporated α-Fe2O3 by the heterogeneous system with H2O2 and UV was investigated by Silva et al. (2009). Niobium was selected, as its ionic ray is compatible with the Fe structure and it shows high reactivity towards H2O2 activation. In addition, Li et al. (2013) reported the photocatalytic degradation of Orange II (OII) using alumina-dispersed α-Fe2O3. This study reported that alumina was an excellent support of α-Fe2O3 nanoparticles and it was more active and stable than silica-supported α-Fe2O3. The alumina-supported catalyst with 25 wt.% Fe2O3 sintered at 400°C had the highest activity according to the result obtained by the authors. It was concluded in the study that alumina-supported FeO with solar light and molecular oxygen could be used to treat wastewater that contains anionic dye.
FeO as a heterogeneous catalyst have been used to improve the efficiency of AOP, particularly Fenton reactions, for dye removal due to their higher activity in a broad range of pH compared to soluble Fe salts. However, FeO catalysts can lose their activities due to the leaching effects of metallic catalysts in acidic medium and this is the major challenge of using FeO catalysts.
6.1.2.2 Zeolite-based catalysts
Zeolites are microporous crystalline aluminosilicates that consist of SiO4 and AlO4 linked to each other by the sharing of oxygen ions. The incorporation of Al into a silica framework (AlO4) makes the framework negatively charged, which can be balanced by extra inorganic or organic cation. Zeolites have internal and external surface areas of 300 to 700 m2/g and cation exchange capacities of up to several milliequivalents per kilogram, which make it a suitable adsorbent and soil modifier for industries (Auerbach et al. 2003). The adsorption capacity and selectivity of zeolites for H2O and/or other molecules depend on the porosity, pore size distribution, and specific surface of zeolites. Elements such as B, Ge, Zn, and P and other transition metals such as Fe, Co, and Mn can be incorporated into the framework of zeolite to change its framework. Such incorporation has led to the development of zeolites with different structures. Cation concentration, sitting, and exchange selectivity vary significantly with Si/Al ratios and play an important role in adsorption, catalysis, and ion exchange applications. Besides, eco-friendliness, cost-effectiveness, and the easy preparation of zeolites attract the attention of researchers (Nezamzadeh-Ejhieh and Banan 2013, Nezamzadeh-Ejhieh and Shirzadi 2014).
Zeolites contain transition metal ions and have been recognized as an efficient heterogeneous catalyst for the degradation of textile industrial effluent (Nezamzadeh-Ejhieh and Banan 2011, Nezamzadeh-Ejhieh and Moazzeni 2013, Apollo et al. 2014, Jamalluddin and Abdullah 2014, Panic and Velickovic 2014). Most of the transition metal/zeolite catalysts are bifunctional (i.e. strong acidic sites are present in the same zeolites). This makes it an ideal catalytic material. Ion exchange and impregnation are two methods generally used for introducing metals into zeolites. One of the limitations of using zeolites as a catalyst is that there is only a small number of zeolites that are suitable as a support in catalysis for synthetic dye removal under acidic condition. Besides, zeolites have selective sites and can only react with a few reactants and product molecules. There are limited researches on using zeolite-based catalysts for the catalytic oxidation of textile dyes, especially synthetic dyes and model wastewater (Table 5). To date, zeolites have been used to degrade heavy metals, adsorbents of benzene, toluene, ethylbenzene, xylenes, phenols, pesticides, herbicides, dyes, and humic acid from natural or industrial wastewater (Aleksić et al. 2010, Nezamzadeh-Ejhieh and Khorsandi 2010a,b, Nezamzadeh-Ejhieh and Salimi 2010, 2011, Nezamzadeh-Ejhieh and Banan 2013, Nezamzadeh-Ejhieh and Bahrami 2014, Nezamzadeh-Ejhieh and Shirzadi 2014).
Fe-containing zeolite is used as a source for ·OH with H2O2. Fe salts are absorbed onto the surface of zeolite, suppressing the reaction between Fe and H2O2 and forming complexes of Fe2+/Fe3+ on the surface of zeolite. The Fe-modified zeolite is easy to prepare due to the ion adsorption and exchange characteristics of zeolite. The loosely bound nature of zeolite allows the exchange of other types of metals in aqueous solution. Fe-exchanged zeolites are found to accelerate the efficiency of Fenton and photo-Fenton oxidation systems. (Tekbaş et al. 2008, Aleksić et al. 2010, Navalon et al. 2010, Gonzalez-Olmos et al. 2012).
Tekbaş et al. (2008) investigated the efficiency of heterogeneous photo-Fenton oxidation of reactive azo dye solutions using Fe-exchanged zeolite catalyst in a quartz batch reactor using artificial UVA as a light source. According to their study, the maximum degradation of 90% was achieved in 60 min at 35°C, pH 5.2, 15 mm H2O2, and 1 g/l of catalyst loading. The stability and reuse of the catalyst were demonstrated in three successive experiments, showing the advantage of heterogeneous catalyst in comparison with homogeneous Fenton reaction. In addition, they also proved that heterogeneous photo-Fenton reaction was successful in the UVA region of solar radiation.
Alver and Metin (2012) investigated the use of modified zeolite with hexamethylenediamine (HMDA) for the adsorption of reactive dyes by batch adsorption experiments through Fenton-like process. The effects of pH, temperature, sorbent dosage, and initial dye concentrations were investigated. It was found that zeolite-modified HMDA improved the absorption capacity of reactive dyes compared to natural zeolites. Natural zeolites were found to have a limited dye removal capacity for anionic dyes. Besides, it was noted that the removal efficiency of modified zeolites for the investigated dyes was independent of the solution pH.
Zeolites have also been successfully used to degrade MB through integrated UV photocatalysis and anaerobic digestion (AD) in up-flow fixed-bed reactors (Apollo et al. 2014). In the study by Apollo et al. (2014), zeolites were applied as the support materials for microorganism and photocatalyst in the bioreactor and photoreactor. In the first part of the experiment, two processes were applied separately, whereas, in the second part, the two processes were combined. It was found that anaerobic degradation performed better in COD reduction (57%), whereas photodegradation performed better in color reduction (70%). However, the integrated process led to high COD and CR efficiency of above 81% and 80%, respectively. The authors reported that UV pretreatment before the AD step improved the biodegradability of MB by three times. As a result, a 2.7 times increase in biogas production was observed for MB that was UV pretreated compared to that without UV pretreatment. Besides, an irradiation time of 120 min was needed for UV photodegradation to achieve the maximum color and COD reduction using a 15 W UV lamp with an energy consumption of 108 kJ. However, the AOP-AD integrated system only consumed 54 kJ of energy to achieve the maximum COD and CR with irradiation duration of 60 min. Therefore, the AOP-AD integration consumed 50% less energy to achieve the same pollutant removal compared to individual AOP.
Besides Fenton-like process and photocatalysis, the effects of zeolites on catalytic ozonation have also been investigated. The influence of the hydroxyl groups of acid-treated natural zeolite on the catalytic ozonation of MB was analyzed by Valdes et al. (2009). The effects of ozonation, adsorption, and integrated system on MB degradation were analyzed. The study revealed that the proton-donating OH groups (Z-OH) of acid-treated zeolites played an important role in the catalytic ozonation of MB by acting as active sites for the adsorption of reacting species. Higher catalytic activity was observed at pH above the PZC. This could be related to the presence of surface hydroxyl groups in the deprotonated form.
In addition, the modification of zeolite surface with quaternary amine hexadecyltrimethyl ammonium bromide (HTAB) to improve the removal efficiency of reactive azo dyes in a zeolite fixed bed was investigated by Benkli et al. (2005). Three different types of dyes [C.I. Reactive Black 5 (RB5), Red 239, and Yellow 176] were used to test the adsorption capacity of the surface of the modified zeolites. The best performance was observed at a flow rate of 0.025 L/min and 3 g/l HTAB dosage. The highest removal percentage of RB5 was observed under optimum conditions. This study also highlighted that a bilayer formation was the most viable packing that enabled maximum removal of the dye.
Several studies have been conducted on the sorbent behavior of zeolites. Zeolites have complex sorption mechanisms due to their porous structure and surface changes. Zeolites have several advantages, including vast availability and low cost, although they are not as effective as clay-based catalysts for dye removal from wastewater.
6.1.2.3 Pillared clay-based catalysts
Clay minerals are abundant on the surface of the Earth and known as attractive materials that have been widely used as adsorbents, catalysts, and catalyst supports (Vicente et al. 2013). They are great adsorbents used for the degradation of contaminants. Clays have a layered structure comprising Si-oxygen tetrahedral and Al-oxygen octahedral (Bergaya and Lagaly 2006, Fanun 2014).
The lack of thermal stability is one of the major limitations of clays. Nevertheless, this can be overcome by modifying clay minerals by acid treatment, thermal treatment, pillaring, and combined thermal and acid treatment. Among them, pillaring is considered the best treatment, as it has a high thermal stability and it does not collapse upon heating at moderate temperature. Modified clays have been receiving increasing attention in recent years due to their chemical catalytic activity. Clays are intercalated with inorganic polyoxocations such as Al3+, Zr4+, Ti4+, and Fe3+ in the interlayer space of the 2:1 clay structure. The polycations in aqueous solutions are intercalated into interlayer of clays by cation exchange and this enlarges the basal spacing. Pillared clay (PILC) has a larger pore size. It is thermally stable and able to provide additional acid sites for the introduction of new oxides. It has been reported that Fe-associated clays are catalytically stable, effective, and more resistant to leaching compared to other Fe species (Carriazo et al. 2003). The preparation of PILCs varies with pH, concentration of metal ions, temperature, and age (Occelli and Robson 1992). Therefore, the differences in the catalytic activity depend on the preparation methods and the composition of the PILCs.
PILC is one of the most widely studied materials among the other groups of microporous materials such as zeolites, silica, carbon, alumina, and titanium. PILCs present higher structural regularity than other classes of adsorbent materials, except zeolites. Extensive reviews on the application of clays in Fenton-like process was published by Garrido-Ramírez et al. (2010) as well as by Navalon et al. (2010). PILCs have been used widely as a catalyst support due to their stability, structure, abundant availability, and cost-effectiveness. The most widely used PILC catalysts include montmorillonites, laponite, bentonites, and saponite.
The use of PILCs in AOPs as a catalyst that helps in the degradation of organic contaminants has been frequently reported in the literatures (Sze Nga Sum et al. 2004, Li et al. 2006, Molina et al. 2006, Ayodele et al. 2012, Galeano et al. 2014). The most frequently used polymeric compounds as pillaring agents are Al, Ti, Cr, and Fe (Bouras et al. 2002, Dvininov et al. 2009, Banković et al. 2012, Chen et al. 2013, Sahel et al. 2014).
The work carried out by Gil et al. (2011) showed that PILCs could be considered as a potential low-cost adsorbent for dye removal from aqueous solutions. In their study, two PILCs were synthesized by the intercalation of aluminum and zirconium solutions and evaluated as adsorbents for the removal of OII and MB. Both clays were found to have the same adsorption capacity when OII was used, whereas the adsorption capacity of Zr-PILC was higher than that of Al-PILC for MB. This showed that dye adsorption was not sensitive to the porosity of the catalyst. The pore size of the two PILCs was between 0.8 and 1.4 nm and this size restricted the diffusion of dye into the pore and caused the differences in reading. Besides, the study reported that an increase in NaCl increased the adsorption capacity of both adsorbents. This was consistent with the accumulation of dye and the presence of electrostatic interactions or other forces between the dyes and PILC surfaces.
Another study has also used Ti-PILC and Ag-doped Ti-PILC for the degradation of methylene dye, which is cationic dye in nature (Sahel et al. 2014). In this study, the authors reported that MB was readily and totally adsorbed on the acid PILC and it disappeared from the solution by cation exchange. However, silver-doped catalysts were found to decrease the activity of Ti-PILC as it had a lower adsorption capacity. Although TiO2-PILC was not as effective as photocatalytic degradation that used TiO2, it could remove pollutants through both adsorption and photocatalysis. In this study, a comparison of degradation efficiency between anionic (Remazol Black) and cationic dyes was also presented. The adsorption efficiency of cationic dye was small in comparison with anionic dye using TiO2-PILC, and reduced efficiency was observed using Ag-doped TiO2-PILC.
Low thermal stability and hydrothermal stability are the main limitations of using PILCs as catalysts. This could be solved by incorporating polycations in PILCs because such incorporation improves the stability, adsorption, and catalytic properties of simple PILC catalysts. Many researches have emphasized using polybasic transition metal oxide PILCs, such as Fe-Ce-Al PILCs and Fe-Cu-Al PILCs (Chen et al. 2013), for dye degradation. A second metal can be introduced in the PILC structure through impregnation and ion exchange (Barrault et al. 2000, Bouras et al. 2002, Damardji et al. 2009a, Banković et al. 2012, Chen et al. 2013, Sahel et al. 2014).
The influence of sorption properties of Fe surfactant-modified pillared montmorillonite (Fe-SMPM) on the Basic Yellow 28 dyes (BY28) and 4-nitrophenol has also been studied at different pH levels in both single component and binary pollutant systems. BY28 was adsorbed onto the modified PILC through hydrophobic interaction and the adsorption increased with pH. An adsorption isotherm of a BY28 system was analyzed and the Freundlich model showed that the result fit well into the model. In their study, the values of Freundlich constants (n>1) suggested that the binding of dyes onto Fe-SMPM was weak and the adsorption was physical. An increase in Freundlich constants (Kf) with pH confirmed that the adsorption capacity was good in acidic medium compared to alkaline environment.
Catalytic wet peroxide oxidation (CWPO) shows an efficient degradation of reactive dyes using PILC catalysts. In the study conducted by Kim and Lee (2004), Al-Cu PILCs were used as catalysts to degrade reactive dyes in a batch reactor through CWPO. Al-Cu PILCs with d001 spacing of approximately 18 Å and surface area of approximately 140 m2/g showed the complete removal of dyes within 20 min at atmospheric pressure and 80°C. This operating condition was milder compared to that of the conventional catalytic wet oxidation process. PILCs were also found to be stable against the leaching of Cu.
Banković et al. (2012) reported the application of Al-Fe-PILCs as heterogeneous catalysts in CWPO of tartrazine azo dye. The pillaring process increased the d001 from 1.28 to ~1.74 nm and SBET values from 123 to 205 m2 g-1. In this study, the influence of Fe and reaction on the removal of pollutant was investigated. The decolorization of tartrazine solutions using Al-Fe PILCs through CWPO was effective with higher Fe3+ content with a dye removal percentage of 97.5% at 75°C after 4 h. It was noted that Al-Fe-PILC catalysts were chemically stable and could be recycled.
PILCs are found to be effective, as they improve the structural characteristics of the clays and increase the accessibility of the reactant molecule. The shape selectivity of clays due to the interlayer and interpillar distances controls the diffusion rate of reactants, reaction intermediates, and products. PILCs are well known as an effective adsorbent for textile dyes in a variety of AOPs, including Fenton, ozonation, and CWPO. The effectiveness of PILCs depends on both the structure of the adsorbent and the targeted pollutant. This is due to the adsorbent-adsorbate interaction that involves both substrate-dye electrostatic interactions and dye-dye hydrophobic interactions. The degradation or mineralization of textile wastewater with PILC catalysts can be an eco-friendly alternative to other advanced treatment techniques. However, PILCs are still used as a batch- or pilot-scale catalyst to date due to their high synthesis cost.
6.1.2.4 Activated carbon
ACs are the most extensively used supports among all the sorbent materials such as zeolite, titanium, and clays for the removal of pollutants from wastewater (Alvárez’ et al. 2009). AC, a highly porous material with amphoteric qualities, is typically employed in wastewater treatment for the removal of organic and inorganic compounds. It is used in both powdered form (PAC) and granular form (GAC). It has attracted researchers due to the important advantages it offers such as low cost, high surface area, high stability, high chemical stability, and easily controlled surface and texture. It is an ideal alternative to other expensive wastewater treatment due to its high efficiency in treating a variety of refractory wastewaters. The types of carbon sorbent and their modes of preparation have significant effects on the adsorption capacity. AC is a promising alternative to homogeneous AOPs for textile wastewater treatment (Lin and Lai 2000, Faria et al. 2005). The excellent adsorption properties of carbon-based supports have been proven for the degradation and decolorization of textile dyes. Many studies have been conducted to test the adsorption of dyes on the surface of AC using AOPs.
Ozonation process, combined with H2O2, UV radiation, metallic ions, or heterogeneous catalysts, has been the subject of intense research recently. The combination of O3 and AC is known as an alternative to degrade colored wastewater (Faria et al. 2004, 2005, Yagub et al. 2014). Heterogeneous catalytic ozonation aims at enhancing the removal of refractory compounds by adsorbing pollutants on the surface of the AC. Ozonation, in the presence of AC, causes the oxidation of organic pollutants by either direct oxidation with O3 or ·OH produced from the interaction between O3 and the AC. The key parameters that affect the activity of AC for the transformation of aqueous O3 into ·OH are carbon dose, O3 dose, carbon type, and pretreatment time of carbon (Sánchez-Polo et al. 2005). ACs with the highest basicity and largest surface areas are the most efficient for organic pollutant removal, as it has the highest adsorption affinity. Besides, the duration of ozonation is found to affect the transformation of O3 into hydroxyl. This suggests that AC does not really act as a catalyst but an initiator and/or a promoter for O3 transformation.
The catalytic activities of AC, CeO2, and ceria-AC composite in terms of the removal efficiency of dyes through catalytic ozonation were studied by Faria et al. (2009). The degradation efficiency of C.I. Acid Blue 113, Reactive Yellow 3, and Reactive Blue 5 was evaluated through the AC-supported ozonation. The performance of the catalytic system was evaluated in terms of decolorization and extent of mineralization. In the study, the ceria-AC composite showed the highest mineralization with 100%, 98%, and 97% TOC removal of C.I. Reactive Blue 5, Acid Blue 113, and Reactive Yellow 3, respectively, after 2 h of reaction time. The experimental studies explained that the adsorption by AC was not efficient for the decolorization of dye solutions, but the combination of AC with O3 enhanced the mineralization of textile effluents. The activity of AC catalysts was not affected by the presence of scavenger species, as it promotes surface oxidation reactions that do not involve ·OH in the solution. Besides, a strong synergic effect between the AC and CeO2 led to high mineralization efficiency. Valdés and Zaror (2006) also mentioned in their study that the effects of radical inhibitors were minimized with the presence of AC, suggesting that carbon surface played a fundamental role in the reaction mechanism. However, the influence of the chemical surface properties of AC on the ozonation of organic contaminants is still unclear.
The heterogeneous Fenton-like process in a continuous packed-bed reactor filled with AC pellets was used to degrade Chicago Sky Blue (CSB; anionic azo dye) wastewater in a study (Mesquita et al. 2012). The AC pellets were impregnated with 7 wt.% Fe. The process yielded dye conversion of 88% and TOC removal of 47% at 50°C, pH 3, Wcat/Q=4.1 g min m/l, and H2O2 feed concentration of 2.25 mm after a consecutive three cycles. Fe leaching was negligible, as it was far below the European Union limits. In addition, Duarte et al. (2013b) investigated the influence of different types of Fe salt impregnated on the commercially available AC for the removal of azo dye OII through heterogeneous Fenton-like process. Fe acetate, Fe sulfate, and Fe nitrate were used as catalysts and each of them contained 7% Fe after pretreatment at 300°C. Three of them showed differences in textural properties such as pore structure and Fe dispersion, which strongly influenced the dye removal adsorption. Among the Fe salts used, ferrous acetate seemed to be the best option based on their activity and stability. Ramirez et al. (2007a,b) investigated the effects of different types of carbon-based catalysts on the degradation and mineralization of nonbiodegradable azo dye OII. The carbon supports were prepared from AC (prepared from olive stone) and carbon aerogel (prepared by the carbonization of an organic resorcinol-formaldehyde polymer). The performance of the catalysts in the Fenton-like oxidation of OII was studied and the effects of the most relevant operating conditions (pH, catalyst concentration, H2O2 concentration, and temperature) were analyzed. In this study, carbon aerogel-based catalyst was found to be more efficient compared to AC. It was clear that dye removal efficiency strongly depends on the types of Fe and structure, compositions, chemical stability of the catalyst used in the system, and the targeted pollutants.
An AC fiber (ACFs) yields great catalytic activity for dye degradation due to its uniform microporous structure and high adsorption ability. ACF-supported 8-hydroxyquinoline ferric has been reported as a heterogeneous Fenton-like catalyst for the efficient removal of dyes (Sun et al. 2014). The developed catalyst exhibits a remarkable catalytic performance across a wide range of pH values (3–9). Moreover, it also shows excellent continued catalytic ability and regeneration capability. This reduces the secondary contamination and cost of the treatment system, as the catalysts can be reused.
Recently, a new magnetic Fenton-like catalyst with GAC as a supporter has been introduced in heterogeneous Fenton-like process for treating dye wastewater (Yang et al. 2014). The study showed that, at optimum conditions ([methyl orange (MO)]0=20 mg/l, temperature=20°C, [NdFeB-AC-FC]0=10 g/l, [H2O2]0=0.6 mmol/l, and pH 3.0), a degradation of 97.1% was achieved after five cycles. The synthesized AC-based catalyst allowed great catalytic activity and was reusable in heterogeneous Fenton-like processes.
In addition, the efficiency of combined O3, PAC, HCO3-, and H2O2 in the decolorization and mineralization of synthetic textile wastewater in a semibatch reactor was investigated (Oguz and Keskinler 2008). The ζ potential values of PAC, ozonated PAC, and ozonated PAC contaminated with intermediates were determined and it was found that PAC acted as an adsorbent in the combined processes. The combined process allowed the COD and CR of 95% and 99%, respectively, at a reaction time of 30 min. It was concluded that O3, PAC, and H2O2 played an important role in oxidizing textile effluents.
It is apparent that ACs are good adsorbents. They are cheap with high surface areas and easily modifiable. The effectiveness of ACs for the removal of a wide range of colored wastewaters has made it an ideal alternative to other treatment options. However, it has several limitations such as regeneration of catalysts, adsorption of secondary molecules, nonselectivity, and ineffectiveness against certain dyes. More studies on the AC-AOP system are required to identify the physical and chemical mechanisms involved.
6.1.2.5 Nanostructured materials
Recently, nanostructured materials have gained enormous attention in wastewater treatment through AOPs (Nezamzadeh-Ejhieh and Banan 2011, Ferroudj et al. 2013, He et al. 2014). Least diffusion resistance, easy accessibility to reactants, and availability of a large number of active sites are among the advantages compared to micro- and macrosized particles in chemical processes especially for wastewater treatment. In most cases, nanoparticles are used as a heterogeneous catalyst in Fenton-like processes.
TiO2 has been the most extensively investigated semiconductor for the photocatalytic oxidation of a variety of dyes (Qi et al. 2005, Khataee et al. 2009, Chebli et al. 2011). It oxidizes organic pollutants through highly reactive oxygen species that is generated by photoexcitation (Fujishima et al. 2000, Daneshvar et al. 2003, Rivas et al. 2012). The heterogeneous photocatalytic decolorization and degradation of Reactive Orange 4 using ferrous sulfate/ferrioxalate, H2O2, and TiO2-P25 nanoparticles was carried out by Selvam et al. (2007). It was observed that the addition of TiO2-P25 nanoparticles improved the photo-Fenton processes. The authors concluded that UV/ferrous/H2O2/TiO2 and UV/ferrioxalate/H2O2/TiO2 were more efficient than the individual photo-Fenton and UV/TiO2-P25 processes. A promising result was observed for the degradation of X-3B dye by immobilized TiO2 photocatalysis coupled with anodic oxidation on boron-doped diamond (BDD) electrode (Zhang et al. 2009a). The nanosized titanium oxides were found to have anatase crystalline structure, uniform nanoparticle distribution, and spherical particle morphology. The oxidants produced on BDD anodes such as H2O2, O3, and peroxodisulfate could raise the quantum efficiency of photocatalytic processes. The experimental results revealed the efficiency of these two AOPs for accelerating dye degradation.
Nanoscale zero-valent (NZVI) Fe could potentially be used in AOPs. A few researchers have successfully applied NZVI/H2O2 to degrade dyes. NZVI offers advantages over zero-valent Fe, as it has higher reactivity due to its small particle size, ability to attach on large particle, large surface area, and possibility to be immobilized in or on suitable solid supports. Shu et al. (2007) investigated the degradation of an azo dye, Acid Black 24 (AB24) solution, using synthesized NZVI Fe particles. The synthesized NZVI particles can effectively decolorize (98.9%) and remove TOC (53.8%) of AB24 dye solution at 30 and 10 min, respectively. The study was conducted at an initial dye concentration of 100 mg/l and an NZVI dosage of 0.3348 g/l. Additionally, the highest removal was 609.4 mg AB24 for each gram of NZVI. The integrated NZVI-H2O2 was employed in an attempt to achieve both complete decolorization and mineralization of the OII solution by Moon et al. (2011). The optimal operating parameters for the degradation of OII by NZVI/H2O2 process based on TOC measurements were NZVI/H2O2=1:16.5, NZVI=200 mg/l (0.357 mm), and pH 3 for the maximum TOC removal of 53% within 60 min. Besides, it should be noted that the prereduction step using NZVI contributed to the overall enhancement in the TOC removal efficiency of OII.
In a study by An et al. (2013), the activity of the nanoparticles could be accelerated by UV radiation. A promising result has been reported for the degradation of methyl violet, Rhodamine B, and phenol using BiFeO3 nanoparticles through visible light photo-Fenton-like catalytic oxidation (An et al. 2013). BiFeO3 nanoparticles are able to catalytically activate H2O2 for the decomposition of organic pollutants under visible light irradiation. In this study, the catalytic ability of BiFeO3 nanoparticles was improved through in situ surface modification by adding a small amount of EDTA (0.4 mm) into the photo-Fenton system. The degradation rate constant of methyl violet was found to increase by 128-fold in the EDTA-BiFeO3-H2O2-photo-Fenton system.
Sheydaei et al. (2014) investigated the decolorization efficiency of textile wastewater using novel γ-FeOOH-GAC nanocomposite by photo-Fenton-like processes in a continuous reactor. The study showed that the decolorization ability of γ-FeOOH-GAC/H2O2/UV process was more effective compared to γ-FeOOH-GAC/H2O2, H2O2/UV, and UV processes at optimum conditions. The optimum operating conditions were flow rate of 1.98 ml min-1, H2O2 concentration of 3.25 mm, and pH 1.98. The synthesized γ-FeOOH-GAC nanocomposites had low leaching rate; therefore, they could be used for a long time. Additionally, maghemite (γ-Fe2O3) nanoparticles and γ-Fe2O3/silica nanocomposite microspheres were evaluated as magnetic heterogeneous Fenton catalysts for the degradation of MB, MO, and paranitrophenol by Ferroudj et al. (2013). The developed nanoparticles maintained a good activity for MO degradation in five repeated experiments despite a relative decrease in the reaction rates.
Besides, carbon nanomaterials are also found to be effective for H2O2 decomposition in aqueous phases. The characteristics of carbon nanomaterials such as low cost, abundant availability and convenient synthesis make them a favorable catalyst support. CNTs with large surface area, mechanical strength, and electrical conductivity have been used in the fabrication of cathode to increase the generation of H2O2 in electrochemical oxidation processes (Khataee et al. 2013a,b). The study conducted by Khataee et al. (2013a,b) proved the efficiency of CNTs cathode for the degradation of C.I. Direct Red 23 by combining photocatalytic and PEF/citrate processes using N-TiO2 nanoparticles under visible light. CNTs were prepared by stabilizing TiO2 nanoparticles on carbon paper using hydrophobic binder (polytetrafluoroethylene). Sol-gel-based method was used to prepare TiO2 nanoparticles. It was found that an efficient immobilization of Fe compounds within the structure of carbon nanomaterials improved the catalytic activity of heterogeneous Fenton-like reaction.
Other nanoparticles such as nanomodified goldmine waste solid have been used as catalysts for the degradation of dye (Huang et al. 2011). Homogeneous Fenton and heterogeneous process were used to evaluate the degradation efficiency of Dispersed Orange 288 in the study conducted by Huang et al. (2011). Fenton process that used nano-Fe-modified goldmine waste solid (68.6%) exhibited the highest degradation efficiency compared to homogenous Fenton process (20.3%).
Nanoparticle is an emerging technology in AOP for wastewater treatment and it has shown great efficiency for the removal of recalcitrant pollutants. The synthesis of more complex nanocomposite materials is given emphasis, as such composite materials could enhance treatment efficiency. Besides, nanoparticles can also be activated by energy-dissipating agents including UV and US. The immobilization of nanoparticles on the support can reduce the environmentally related problem, as post-treatment is not necessary. Nanomaterials such as CNT need to be recycled before being discharged to the environment, as the associated health risk is still unclear.
6.2 Energy-dissipating agents
6.2.1 Ultrasound
Since 1990, US has gained attention in environmental engineering because it shows great potential in degrading organic pollutants from wastewaters (Mason 1999, Liang et al. 2007, Eren 2012). The chemical process in which US irradiation is involved is called sonochemistry.
·OH is generated in aqueous solutions when exposed to ultrasonic waves followed by subsequent oxidation of pollutants. US generates chemical effects though different physical mechanisms, and cavitation is the most important nonlinear acoustic process for sonochemistry. It can be generated through two different ways: (i) hydrodynamic cavitation (when large pressure differentials are applied in flowing liquid) and (ii) acoustical cavitation (through which electrochemical transducer is in contact with the fluid; Lewinsky 2007). Cavitational phenomena including the formation, growth, and collapse of acoustic bubbles affect the oxidization of organic pollutants (Mason 1999). When liquid is sonicated, dissolved gas molecules are entrapped by microbubbles that grow and expand upon the refraction of acoustic cycles. Then, cavitation bubbles are formed. These bubbles continue to grow in size until the maximum of the negative pressure and then release extreme temperature upon adiabatic collapse. The organic pollutants present in wastewater are directly decomposed or indirectly oxidized by the formed ·OH during the momentary collapse of cavitation bubbles (Lin et al. 2008). High temperatures (2000–5000 K) and pressures (1000–1800 atm) induced by cavitations in aqueous solutions lead to the thermal dissociation of H2O into reactive ·OH and hydrogen radical [Equation (27)].
Sonochemical effects take place at the gas-liquid interface due to the oxidation of organic molecules by ·OH and, in certain cases, in the bulk solution or during pyrolytic decomposition inside the bubbles. Acoustic cavitation is dominant for organic pollutant degradation, as it generates hydromechanical forces and pyrolytic reaction. Hydrophilic and nonvolatile compounds such as dyes are mainly degraded through OH-mediated reactions in the bulk solution and at the gas-liquid interface, whereas hydrophobic and volatile species degrade thermally in the interface region of the collapsing bubbles. Researchers have found that high frequencies are usually effective in degrading dyes compared to low- and medium-frequency US. In recent years, US has been extensively used as an AOP for wastewater treatment. A number of reviews that cover the utilization of US for wastewater treatment through ozonation, photolysis, photocatalysis, electrochemical oxidation, and Fenton AOPs for organic pollutant degradation are available (Gogate 2008, Mahamuni and Adewuyi 2010, Pang et al. 2011, Eren 2012, Bagal and Gogate 2014). Sonochemical reaction can disintegrate the molecular structure of different types of dyes beyond the audible range (20–1000 kHz) (Adewuyi 2001, Mason and Lorimer 2003). There are a few studies on the ultrasonic degradation of textile effluents or wastewater.
Although US is effective in degrading pollutants, total mineralization using US alone is almost impossible, especially for recalcitrant wastewater. To increase the decomposition efficiency and reduce the time required for degrading recalcitrant pollutants, combinations of US with different oxidants such as O3 or H2O2 have been widely investigated by researchers. Since dyes are nonvolatile, oxidative ·OH reactions in the bulk solution are expected to be the major contributing factor for pollutant degradation. US can enhance the mass transfer rates of solutes and improve the surface properties of solid particles. Recent researches focus on hybrid techniques such as US/O3, US/H2O2, US/UV, US/Fenton, UV/US/TiO2, UV/US/ZnO, and UV/US/O3 to enhance the degree of mineralization of pollutants (Wu 2007). Ultrasonic irradiation combined with heterogeneous catalysts is the most commonly used for dye degradation (Lin et al. 2008, Zhang et al. 2009b, Eren 2012).
Combinations of US/Fenton reagent have been proven the simplest way to generate ·OH for textile wastewater treatment (Lin et al. 2008, Jamalluddin and Abdullah 2014). Fenton reagent along with US decolorized RB5 completely at the optimum experimental conditions of pH 3, 8.82 mmol/l H2O2, and 0.5 mm Fe2+ (Saravanan and Sivasankar 2014). Recent studies on Fenton process integrated with US have proven the efficiency of such process to decolorize colored solution. Its synergy factors are higher compared to US and Fenton alone. However, it should be noted that the use of US in AOPs is not energy efficient, whereas the use of a catalyst is found to improve energy consumption through synergistic effects. It is should be noted that the smaller the particle size of the catalyst is, the greater the synergy effect it has. For example, TiO2-combined US is believed to enhance the number of nucleation sites, which increase the generation of ·OH. Electron-hole pairs in the semiconductor catalyst are generated through thermal energy, leading to the generation of additional ·OH. Additionally, studies have been conducted on the effects of Ag/TiO2 nanocomposites in US combined with UV on the decolorization of RR2 (Wu 2009). In the study, it was found that, after 120 min of reaction, the TOC degradation efficiencies of UV/US/TiO2 (63%) exceeded that of the UV/TiO2 system (47%). A similar result was reported by Kritikos et al. (2007). It is noted that the simultaneous application of UV and US irradiation resulted in the highest decolorization efficiency compared to that achieved by photocatalysis and sonolysis separately.
It is clear that the advantages of US such as high penetration, high degradation efficiency, high energy conservation, and zero generation of secondary pollutants are the main reasons it has been used widely for wastewater treatment, including dye degradation. Based on the literature study, sonolysis may not be suitable as a single process, but coupling it with other AOP techniques results in higher degradation and mineralization efficiency (Liang et al. 2007, Gogate 2008, Mahamuni and Adewuyi 2010, Taha et al. 2014, Khataee et al. 2015a,b,c). However, the synergic effects are not clearly known. It is necessary to conduct more researches to optimize the operating parameters by focusing on the kinetic studies, as the kinetics of a particular system has an impact on the operating cost of a treatment system.
6.2.2 UV radiation
UV radiation-based technologies have emerged as an important wastewater treatment method for the past three decades. Disinfection is the primary purpose of using UV radiation in H2O treatment (Oppenländer 2007). The effectiveness and economic feasibility have made UV disinfection viable for industrial wastewater treatment. UV system is a chemical-free H2O treatment system with a low maintenance and operational requirement. It has first been used in France for drinking H2O treatment followed by groundwater treatment since 1906. At present, UV has gained attention for the removal of refractory organic pollutants from wastewaters.
UV electromagnetic radiation covers the wavelength range between 10 and 400 nm, which can be subdivided into few categories based on the wavelength range. UV radiation can be distinguished between UV-A (400–315 nm), UV-B (315–280 nm), UV-C (280–200 nm), and vacuum-UV (200–100 nm; Masschelein and Rice 2002). Among them, UV-C is usually used for H2O disinfection by researchers. Among the commercially available lamps, mercury lamps have been widely used to emit UV radiation, as they have very low ionization energy to enable the avalanche effect. Three types of mercury lamps are commercially available including low-, medium-, and high-pressure lamps. A low-pressure mercury lamp is commonly used for UV radiation. The degradation of H2O contaminants could be achieved through direct photolysis at 254 nm or indirect oxidation by ·OH at irradiation of 185 nm. Oxidation via ·OH is influenced by the presence of organic and inorganic pollutants. In addition, the last decade has witnessed a growing research focus on the application of excimer and excilamp lamps for disinfection and degradation purposes.
UV demonstrates its versatility and significance for current developments related to AOPs. Most of the photoinitiated processes are radical reactions. In AOPs, ·OH is the reactive species mainly responsible for the photo-oxidation of organic compounds. UV is widely used as agents for the cleavage of chemical bonding and the removal of refractory compounds in AOPs. The use of UV in AOPs for wastewater treatment is on a rising trend. AOP with UV irradiation has better efficiency in removing toxic compounds. UV irradiation can be used as a single technique or in combination with other AOPs to offer both ecological and economic advantages. The combination of UV with other AOPs such as homogeneous and heterogeneous Fenton-based processes, catalytic ozonation, electrochemical oxidation, sonolysis, H2O2, and heterogeneous photocatalytic oxidation produces a highly effective result, which could not be achieved if they are used separately (Saien et al. 2011, Saekow et al. 2012).
Direct photolysis and ·OH reactions are the two most important pathways to degrade contaminants in UV-AOP systems [refer to Equations (28) and (29); Zhang et al. 2013].
The residual effect of UV radiation is highly dependent on the constituents of the untreated effluents. The presence of organic compounds that may act as a photosensitizer influences the residual effects of UV radiation. Besides, photocatalytic reaction rate is very much dependent on UV photon flux. An increase in photon flux causes the rate constant to change from first-order to half-order. It is because change in UV light intensity accelerates the Fe redox cycling rate and increases the generation of ·OH. In heterogeneous photocatalytic oxidation, electron-hole pair recombination is enhanced with increased UV photon flux and the addition of H2O2 increases the reaction rate and makes the relationship between UV intensity and degradation rate closer to first-order reaction (Zhang et al. 2013).
In addition, high UV dose in combination with other oxidants may help degrade harmful by-products from the reaction and the parental compounds (Shu and Chang 2005a, Wols and Hofman-Caris 2012). In Fenton process, UV light irradiation increases the rate of organic pollutant degradation by forming additional hydroxyls and recycling ferrous catalyst by reducing Fe3+ (Vujević et al. 2010). Besides, illuminating pollutants with high-intensity UV in O3/TiO2 treatment also increase photocurrent density and energy conversion efficiency of the system (Saekow et al. 2012). This is because high-intensity UV reduces the charge transfer resistance and increases the absorption of dye on the surface of TiO2 catalysts.
When AOPs are applied individually, they are expensive and energy intensive. Therefore, the integration of UV with other AOPs should be applied to reduce the cost of treatment and increase the treatment efficiency at the same time. However, UV shows some limitations, as the concentration and turbidity of the targeted compounds affect the distribution of radiation in the treatment system. It is important to design an effective photoreactor by considering the key parameters that affect AOPs, as discussed earlier.
7 Challenges for the future
7.1 H2O2
The main drawback of AOPs is the high operating cost associated with the use of costly chemicals, especially H2O2 (Oller et al. 2011). The in situ production of H2O2 has gained great interest from researchers recently, but more studies are needed on the efficient in situ production of H2O2. H2O2 consumption can be reduced through continuous feed throughout the reaction and proper optimization of AOPs.
7.2 Toxicity assessment
The environmental impact of the treatment system in terms of toxicity is one of the most important aspects that need to be taken into consideration to evaluate AOPs. The treated effluent should be less harmful or toxic compared to the initial effluent. This is because some products formed during the treatment process may magnify the harmful effect of a contaminant; for example, long-lived intermediates could be more toxic than the original dye.
7.3 Reactor design
AOP reactors are rarely found in the literatures, as designing an efficient large-scale reactor is challenging. A few factors need to be carefully considered while designing a reactor, so that there is uniform irradiation on solutions at the incident light intensity. Besides, mass transfer problems occur in O3-based systems.
7.4 Presence of ions in wastewater
The presence of anions and cations may affect the degradation processes via reaction with ·OH ions and absorption of UV light. CO3-, HCO3-, CI-, SO24-, PO43-, and NO3-, which have significant impact on contaminant degradation processes, are examples of ions that are present in wastewaters. Therefore, more studies need to be conducted to analyze the effects of anions and cations on AOPs.
7.5 Limitation of catalyst
It is necessary to develop a catalyst with high activity and low cost because expensive catalysts are not practical for industrial application. Besides, leaching of catalytically active species by intermediates is the main limitation of a heterogeneous catalyst that needs to be overcome for the application of the catalyst in real wastewater. Catalyst deactivation may occur due to catalyst specific surface area, poisoning of catalytic agents by intermediates formed, surface deposition, and strong adsorption on a polymeric carbon layer. In addition, the catalysts developed should also be easily recoverable for further reuse.
7.6 Formation of by-products
AOPs may form by-products with higher polarity and H2O solubility than the parent compounds. These by-products may be toxic to human beings and aquatic lives. Therefore, further studies should be conducted to investigate the efficiency of combining a variety of catalysts with AOPs.
7.7 Industrial-scale applications
Although AOPs are highly successful at laboratory scale, more studies are needed to evaluate their applicability at industrial scale. Economic viability and efficiency of AOPs are major factors that limit their industrial application. Among AOPs, ozonation and UV/H2O2 have been used widely at industrial scale for drinking H2O treatment. Table 7 summarizes the existing AOP treatment systems at industrial scale. Based on Table 7, it is clear that AOPs are widely used for drinking and groundwater treatment but rarely used in wastewater treatment.
Application of AOPs in industrial scale (Sarathy and Mohseni 2010, Audenaert et al. 2011).
| Treatment system | Company | Pollutant |
|---|---|---|
| Ozonation | Cadillac Motor Car Division of General Motors Corporation, Detroit, MI, USA | Cyanide removal |
| Ozonation | Bell telephone laboratories in 1973 | Control of bacteria |
| O3/UV | Tinker Air Force Base, OK, USA | Metal complexed cyanides and refractory organics |
| O3/GAC | ARCO Products Company, Richmond, CA, USA | Petroleum industry |
| UV/H2O2 system | Milan Army Ammunition Plant, Milan, TN, USA | To treat holding ponds contaminated with explosive compounds |
| UV/H2O | Air Force and EPA Demo, Edwards AFB, CA | Organic contaminants |
| UV/H2O | U.S. Navy Site, NJ, USA | Groundwater (organic contaminants) |
| UV/H2O | Winthrop Superfund Site, ME, USA | Groundwater (pretreatment for Fe) |
| UV/H2O | Milan AAP, Milan, TN, USA | Groundwater (explosives) |
| UV/H2O2 | PWN H2O treatment plant in North Holland | Treat micropolultants (herbicides, pesticides, etc.) in the drinking H2O |
| UV/O3/H2O2 | Bofors Nobel Superfund Site, located near Muskegon, MI, USA | Decomposition of hazardous wastewaters containing benzene, toluene, chlorobenzene, tetrachloroethane, and benzidine |
| UV/H2O2 | Trojan Technologies has installed such a system in Cornwall, ON, Canada | Drinking H2O (to treat taste and odor compounds) |
| UV/H2O2 | San Jose, CA, USA | Groundwater (halogenated hydrocarbons, VOCs, pesticides, PCBs, TCE, 1,1-DCA, and 1,1,1-TCA) |
| UV/H2O2/O3 | Kansas City Plant, MO, USA | VOCs |
| Fenton system | Chemical factory in South Poland | Removal of color, mineralization (COD), and toxicity removal |
| Fenton system | Specialty Chemical Manufacturer, LA, USA | Phenol compounds |
| Fenton system | Refinery, Southeast USA | Pretreatment |
| Fenton system | Wood Treating Facility, Western USA | Phenols, naphthols, and cresols |
| Fenton system | Aircraft Painting Stripping and Maintenance Facility, Midwest USA | Reduce toxic organics before discharge to the municipal wastewater collection system |
| Fenton system | Chemical Plant, AL, USA | Phenol |
VOC, Volatile organic compounds.
It should be noted that the industrial application of AOPs, especially Fenton-based processes for wastewater treatment, has not been much explored. Most of the previous studies have been conducted to evaluate the degradation performance using simulated wastewater. Limited research has been conducted on the effectiveness of AOPs for real effluent treatment. The degradation mechanisms vary with contaminants. Therefore, comparison between AOPs should be made in terms of removal efficiency, operation safety, manageability, applicability, technical effectiveness for real wastewater, necessity of post-treatment, secondary pollution prevention, toxicity reduction, and economic viability.
There are limited studies on the cost estimation of different AOPs. The cost of different AOPs includes capital, operating, and maintenance costs. An overall kinetic model is required to design AOPs and determine the most efficient or economic operating regions. Along with the full-scale kinetic models, such cost is dependent on operating parameters, such as H2O2 concentration, initial pollutant content, irradiation of light, and catalyst concentration. Overall, more efforts are needed to establish the most efficient and cost-effective treatment methods for industrial-scale systems.
8 Conclusion
This manuscript provides a critical review on potential AOPs for textile wastewater treatment. It is estimated that approximately 15% of the dyestuffs are lost in industrial effluents during manufacturing processes. The conventional treatment methods are inefficient to remove refractory compounds in the colored effluent.
AOP is a promising pretreatment and advanced treatment process for the degradation and mineralization of refractory or recalcitrant pollutants in various types of textile wastewaters. The most extensively studied conventional AOPs include O3-based processes, Fenton and Fenton-like, photo-Fenton and photo-Fenton-like, heterogeneous photocatalytic oxidation, and electrochemical oxidation process. Their efficiency in the removal of recalcitrant compounds from textile wastewaters has been studied. In comparison with the other conventional AOPs, Fenton based are found to be a better method for producing extremely reactive ·OH that is responsible for the degradation of pollutants, as no special reactants or equipment are required compared to other AOPs such as ozonation. Fenton reagents are readily available, easy to store, safe to handle, and nonthreatening to the environment. The process can also be performed at ambient temperature and pressure without a complex system. Besides, electrochemical oxidation is also found to be one of the emerging methods for wastewater treatment, but more work is needed to explore the efficiency of the treatment system.
The key parameters that affect the efficiency of AOP treatment systems have been reported in the literatures, and particular attention is given to the effect of these key operating parameters on process performance. The key parameters include the initial concentration of oxidants, pollutant concentration, catalyst load, pH, and irradiation of light. Although AOPs have been recognized as an efficient chemical oxidation method for wastewater treatment, there is a need for additional research to enhance the treatment efficiency. This manuscript explains some of the major drawbacks of AOPs, including sludge formation, mass transfer limitation, pH dependence, and high treatment cost.
This paper also discusses the current enhancement of conventional AOPs using heterogeneous catalysts and energy-dissipating agents to overcome the limitations of conventional AOPs. Heterogeneous materials such as FeO minerals, zeolite-based catalysts, PILC, ACs, and nanosized materials exhibit high catalytic activity in the oxidation processes compared to homogeneous catalysts. The usage of heterogeneous catalysts is favorable because the complete mineralization of organic pollutants can normally be achieved with the easy separation of the catalysts and there is no generation of secondary pollutants. Further studies are required to further explore the potential of different types of systems to enhance the efficiency of pollutant removal with minimal treatment cost. Combined processes have thus become a promising alternative to replace conventional AOPs for the mineralization of recalcitrant contaminates. The combined processes offer more advantages over conventional AOPs by being cost-effective, nontoxic, eco-friendly, and biocompatible.
However, further studies are required to identify more cost-effective AOPs. In addition, more studies should be conducted at a large scale instead of batch scale to exhibit the industrial applicability of AOPs. Because the relationship between pollutants and their reaction pathways makes it difficult to extrapolate the results obtained from one pollutant to another, more studies that use real effluents are needed. Real wastewater should be treated using integrated process based on the thermodynamics and reaction kinetic studies, as they are important for cost estimation.
In addition, mechanistic and mathematical modeling can be used to optimize and define the relationship between parental compounds and reactants used in the treatment systems. Additionally, hybrid AOP methods may also offer better removal efficiency of refractory compounds when compared to individual AOPs, as some of the drawbacks of other AOPs can be eliminated. Normally, combinations of different AOPs enhance the generation of free radicals, which increase the reaction rate.
Overall, it can be concluded that AOPs can potentially become a powerful technique if the complexity of designing and optimizing the system is addressed and solved. Besides, all the factors that contribute to enhancement of degradation efficiency should be carefully explored and evaluated. There should be more studies on the application of AOPs at industrial scale.
About the authors
Archina Buthiyappan graduated from the University Technology of Malaysia with a bachelor’s degree in industrial chemistry in 2008 and a Master’s degree in forensic science in 2010. She joined the University of Malaya (Malaysia) as a doctoral candidate in 2012. Her research focus includes the application of various types of advanced oxidation processes such as Fenton, photo-Fenton, and electro-Fenton to treat real textile effluents.
Abdul Raman Abdul Aziz completed his PhD in the area of three-phase mixing. Currently, he is a professor and holds the position of Deputy Dean at the Faculty of Engineering, University of Malaya (Malaysia). His research interests are in advanced wastewater treatment and mixing in stirred vessels. Before joining the University of Malaya, he worked in the oil and gas and food industries from 1989 to 1993. He is also active in consultancy projects and is currently supervising many PhD candidates. To date, he has published more than 100 papers in journals and conference proceedings both locally and internationally. He is also a member of professional and learned societies, such as the Institution of Chemical Engineers (UK) and the Institution of Engineers Malaysia.
Wan Mohd Ashri Wan Daud is a professor of chemical engineering at the University of Malaya (Malaysia). He earned his bachelor’s degree in chemical engineering in 1991 from Leeds University (Leeds, UK) and Master’s degree in chemical engineering in 1993 from the University of Sheffield (Sheffield, UK). In 1996, he obtained his PhD in chemical engineering from the University of Sheffield. His research fields include fuel cells, energy, biomass conversion and the synthesis of catalyst materials, catalysis, zeolites, polymerization process, separation processes (adsorption, activated carbon, and carbon molecular sieve), ordered mesoporous materials, and hydrogen storage materials. Professor Daud has published approximately 90 research papers.
Acknowledgments
The authors are grateful for the University of Malaya High Impact Research Grant (HIR-MOHE-D000037-16001) from the Ministry of Higher Education Malaysia and University of Malaya Postgraduate Research Fund, which financially supported this work.
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