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International Journal of Photoenergy
Volume 2014 (2014), Article ID 821674, 21 pages
Recent Developments in Homogeneous Advanced Oxidation Processes for Water and Wastewater Treatment
1Department of Civil and Environmental Engineering, Water & Environmental Technology (WET) Center, Temple University, Philadelphia, PA 19122, USA
2Laboratory of Green Chemistry, Faculty of Technology, Lappeenranta University of Technology, Patteristonkatu 1, 50100 Mikkeli, Finland
3Department of Environmental Engineering and Science, Feng Chia University, No. 100 Wenhwa Road, Seatwen District 407, Taichung, Taiwan
4Department of Chemistry, Annamalai University, Annamalai Nagar 608002, India
Received 31 May 2013; Accepted 21 October 2013; Published 26 February 2014
Academic Editor: Vincenzo Augugliaro
Copyright © 2014 M. Muruganandham et al. This is an open access article distributed under the Creative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.
This paper reports on recent developments in homogeneous Advanced Oxidation Processes (AOPs) for the treatment of water and wastewater. It has already been established that AOPs are very efficient compared to conventional treatment methods for degradation and mineralization of recalcitrant pollutants present in water and wastewater. AOPs generate a powerful oxidizing agent, hydroxyl radical, which can react with most of the pollutants present in wastewater. Therefore, it is important to discuss recent developments in AOPs. The homogeneous AOPs such as O3, UV/O3, UV/O3/H2O2, and UV/H2O2, Fe2+/H2O2, UV/Fe2+/H2O2 on the degradation of pollutants are discussed in this paper. The influence on the process efficiency of various experimental parameters such as solution pH, temperature, oxidant concentration, and the dosage of the light source is discussed. A list of contaminants used for degradation by various AOPs and the experimental conditions used for the treatment are discussed in detail.
Wastewater is water that contains various pollutants, which means it cannot be used like pure water and should not be disposed of in a manner dangerous to humans, living organisms, and the environment. Water pollution has a serious impact on all living creatures, adversely affecting water use for drinking, household needs, recreation, fishing, transportation, and commerce. It has been estimated that the total global volume of wastewater produced in 1995 was in excess of 1,500 km3 . On July 28, 2010, the United Nations General Assembly declared safe and clean drinking water and sanitation a human right essential to the full enjoyment of life and all other human rights . It is a concern that nearly 900 million people in the world do not have access to safe drinking water. Approximately 1.5 million children under five die every year as a result of diseases linked to a lack of access to water and sanitation as indicated by World Health Organization (WHO) . It was estimated that about 1.8 million deaths annually are due to lack of access to safe drinking water and poor sanitation.
In the past, economically viable chlorination has been used for water treatment. Yet the potentially adverse health effects of the by-products formed, together with raised drinking water standards, have led researchers to search for effective and economical alternatives to chlorinating drinking water [4, 5]. Various wastewater treatment processes have been tried using physical, chemical, and biological methods [6–12]. Some of these methods have disadvantages, however, and cannot be applied for large scale treatment. For example, one drawback of precipitation methods is sludge formation. Chemical coagulation and flocculation use a large amount of chemicals and the generated sludge may contain hazardous materials, so sludge disposal remains a problem. Adsorption techniques have been used widely for the removal of various water and wastewater pollutants. Their disadvantage is that the pollutants may only transfer to the adsorbent, which needs to be regenerated regularly, resulting in additional costs. Membrane technologies such as ultrafiltration, nanofiltration, and reverse osmosis have been used for the full scale treatment and reuse of water and chemicals. Yet these methods have several operational difficulties in addition to high capital costs. Thus physical methods may not be suitable for the complete removal of pollutants from the environment. Similarly, two different basic biological wastewater treatment methods have been employed: aerobic and anaerobic treatments. These methods also do not completely remove the high concentration of pollutants present in wastewater. Other biological methods involve cost-effectiveness or operational difficulties, making biological means unsuitable for wastewater treatment.
Among the chemical methods, oxidation is efficient and applicable to large scale wastewater treatment. Generally air, oxygen, ozone, and oxidants such as NaOCl and H2O2 are used for chemical treatment. The oxidation potential of some of the oxidants is listed in Table 1. The basic chemical oxidation process with air and oxygen also occurs in nature, but it is no longer sufficient for highly polluted wastewater. Therefore there is a significant need to develop a wastewater treatment process which can remove the pollutants effectively by a simple method.
Advanced oxidation processes (AOPs) for wastewater treatment have received a great deal of attention in recent years. AOPs generate the highly reactive hydroxyl radical (•OH) to degrade the recalcitrant chemicals present in wastewater [13–15]. These OH radicals attack the most organic molecules rapidly and nonselectively. The versatility of AOPs is also enhanced by the fact that they offer various alternative methods of hydroxyl radical production, thus allowing a better compliance with specific treatment requirements. The eco-friendly end product is the special feature of these AOPs, which are more efficient as they are capable of mineralizing a wide range of organic pollutants. Interestingly, AOPs can make use of solar energy rather than artificial light sources. The latter rely on high electrical power, which is costly and hazardous.
AOPs such as ozonation (O3), ozone combined with hydrogen peroxide (O3/H2O2) and UV irradiation (O3/UV) or both (O3/H2O2/UV), ozone combined with catalysts (O3/catalysts), UV/H2O2, Fenton and photo-Fenton processes (Fe2+/H2O2 and Fe2+/H2O2/UV), and the ultrasonic process and photocatalysis have been successfully used for wastewater treatment [16–23]. This review reports on recent advances in the aforementioned AOPs for water and wastewater treatment. The authors discuss the principle of hydroxyl radical generation from each AOP, the influence of various experimental parameters, and their consequences for the treatment process.
1.1. Ozone Based Advanced Oxidation Processes
(i) Ozonation. Ozone is an environmentally friendly oxidant since it decomposes into oxygen without producing self-derived by-products in the oxidation reaction. It is widely used in the purification of drinking water, the treatment of wastewater and process water, the sterilization of water in artificial pools, and so forth. In an ozonation process, two possible oxidizing actions may be considered. The first or direct method involves the reaction between ozone dissolved compounds. The second is known as the radical method because of the reactions between the hydroxyl radicals generated in ozone decomposition and the dissolved compounds . Some oxidation products are refractory to further oxidative conversion by means of ozone, thus preventing a complete abatement of TOC. Yet the high energy cost of direct ozonation limits many practical applications. To increase the efficiency of the ozonation process, the ozone is combined with H2O2 and UV light, which is expected to increase the removal rate substantially by producing more hydroxyl radicals in the treatment system.
(ii) O3/H2O2 Process. Hydroxyl radicals are generated by a radical chain mechanism through the interaction between ozone and H2O2 as shown in (1). Degradation is facilitated by both ozone and hydroxyl radical:
(iii) O3/UV Process. Hydroxyl radicals are generated in the O3/UV process by photolysis of ozone in the presence of water as shown in the following:
(iv) O3/UV/H2O2. This combined process may generate hydroxyl radicals in different ways as mentioned in (1)-(2). It is considered to be the most effective treatment process for highly polluted effluents.
Wastewater was treated using the Fenton process or homogeneous AOP employing iron salt with hydrogen peroxide. The combination of Fenton’s reagent with UV light is called a photo-Fenton reaction. UV light irradiation enhances the efficiency of the Fenton process. The hydroxyl radical generated in the Fenton process is due to the iron catalysed decomposition of H2O2 as shown in the following: In addition to the above reaction the formation of hydroxyl radical also occurs by the following reactions in the photo-Fenton process shown in the following: This is also attributed to the decomposition of the photoactive Fe(OH)2+ which leads to the addition of the HO• radicals: A considerable increase in oxidation power is observed mainly due to the photoreduction of Fe(III) to Fe(II), which can react with H2O2, establishing a cycle: Among various AOPs, Fenton’s reagent (H2O2/Fe2+) is one of the most effective methods of organic pollutant oxidation. The advantage of Fenton’s reagent is that no energy input is necessary to activate hydrogen peroxide. These processes are economic and can be operated and maintained easily.
1.2. UV/H2O2 Process
The UV/H2O2 process is a homogeneous advanced oxidation process employing hydrogen peroxide with UV light. Hydrogen peroxide requires activation by an external source such as UV light and the photolysis of hydrogen peroxide generates the effective oxidizing species hydroxyl radical (•OH). The rate of photolysis of H2O2 depends directly on the incident power or intensity. The hydrogen peroxide decomposition quantum yield is 0.5 at UV (254 nm) irradiation. Solar light could also be used as a radiation source but the rate of photolysis may be low compared to UV light. In this process the dosage needs to be optimized, however, since excess H2O2 may scavenge hydroxyl radical.
1.3. Heterogeneous AOPs
1.3.1. Catalytic Ozonation Process
Heterogeneous catalytic ozonation is a novel type of AOP that combines ozone with the adsorptive and oxidative properties of solid phase catalysts to decompose pollutants at room temperature. Catalytic ozone decomposition at room temperature is advantageous compared to thermal decomposition in terms of energy conservation since it does not require large volumes of air to be heated. It is therefore a promising advanced oxidation technology for water treatment.
Heterogeneous photocatalysis through illumination by UV or visible light on a semiconductor surface generates hydroxyl radicals. The photocatalyst can be used successfully for the effective treatment of pollutants in water and wastewater.
2. Ozone Based AOPs
As noted above, ozone reacts with various organic and inorganic compounds in an aqueous solution, either by direct reaction of molecular ozone or through a radical mechanism involving hydroxyl radical induced by the ozone decomposition. Figure 1 shows the experimental setup of the ozonation process. This process is strongly influenced by a number of experimental parameters such as solution pH, influent ozone dosage rate, and temperature. The primary reactions initiated by ozone in water are strongly pH dependent. Ozone reacts with organic substrate at low pH as a molecular form, but at high pH it decomposes before reacting with the substrate. Ozone decomposition is catalyzed by hydroxide ions and proceeds more rapidly with increasing pH, eventually to produce hydroxyl radicals. The influence of solution pH on ozonation process efficiency has been observed in a number of studies. For example, Jung et al. investigated the effect of pH on the ozonation of ampicillin from pH 5 to 9, concluding that higher pH conditions are necessary for effective removal . They also discussed how changing pH influences the charge of some specific functional groups on the ozonation process. Can and Gurol investigated the effect of solution pH on the ozonation of humic substances. They found that rapid ozone decomposition was caused by the interaction of ozone with the humic substance, which eventually yielded hydroxyl radical . They further noted that increasing humic substance concentration facilitates fast ozone decomposition into hydroxyl radical. Similarly, the influence of solution pH and temperature on the ozonation of six dichlorophenols was investigated by Qiu et al. . They revealed that the changing solution pH was strongly influenced the decomposition and the rate was increased by raising the hydroxyl radical concentration from acidic to alkaline pH .
Although hydroxyl radical formation is highly favourable to produce more •OH radicals by ozone self-decomposition at pH 10, a portion of carbonate or bicarbonate ion formation could play a key scavenging role in trapping •OH radicals, appreciably decreasing the degradation rate. Wu et al. found that 2-propanol degradation decreases at pH 10 and suggested bicarbonate formation as the possible reason for the decreasing degradation rate at this pH . Other studies reached quite different results. Moussavi and Mahmoudi noted a higher removal rate of Reactive Red 198 azo dye in an ozonation process at pH 10 . Interestingly, Begum and Gautam noted that as the pH increased from 9 to 12 in the ozonation process the endosulfan and lindane removal rate also increased . In contrast to the above results other authors noted that the oxidation rate is relatively independent of solution pH values . Hong and Zeng found that the rates of pentachlorophenol decomposition were very similar between pH 7 and 12, indicating then negligible influence of pH values . These results clearly showed that the nature of pollutants being used for the ozonation process played an important role besides the favourable hydroxyl radical formation at higher pH. Based on the above discussion it is concluded that the influence of pH on the ozonation process needs to be optimized.
Several investigations were conducted into the effect of temperature on the ozonation process. Changing the temperature generally influences the ozonation process in two ways. Firstly, when the temperature increases, the solubility of ozone may decrease, since Henry’s law coefficient of ozone increases with rising temperatures. Secondly, raising the temperature increases the activation energy which may positively assist the ozonation process. Muruganandham et al. noted that N-methyl pyrrolidone (NMP) mineralization was substantially increased when the ozonation temperature rose from 5 to 50°C . They also concluded that the increasing removal rate due to the higher reaction temperature is not balanced by the lower solubility of ozone. Similar results were noted in other ozonation studies [68, 73–76]. Some researchers found, however, that increasing temperature in the ozonation process decreases the removal rate by decreasing the ozone solubility [77, 78]. Interestingly, Ku et al. found that the reaction rates of phorate decomposition were relatively independent of solution temperatures and pH values . Yet some mineralization formation of products such as phosphate and carbonate was increased significantly with raised solution temperature.
Another important experimental parameter influencing ozonation process efficiency is influent ozone dosage. Treatment cost increases with a higher applied ozone dose, so it is necessary to optimize this dosage. For semibatch experiments, increasing the ozone dosage will enhance the mass transfer rate of ozone from the gas phase to the liquid phase, which is expected to enhance the degradation rate appreciably. As the ozone concentration in the liquid phase is saturated, however, ozone mass transfer is limited at a very high ozone dosage . Many authors investigated how the influent ozone dosage affects the degradation rate in the ozonation process within different experimental parameters. Muruganandham et al. reported that the optimal ozone dosage for NMP mineralization is 18.4 mg min−1 . Moreover, an ozone dosage of 27.6 mg min−1 was noted as optimum for the degradation of dimethyl sulphoxide (DMSO) . Begum and Gautam reported an optimum ozone dosage of 57 mg min−1 for endosulfan and lindane degradation although a higher ozone dosage slightly increased the endosulfan decomposition . Yet other studies reported a linear increase in removal efficiencies with ozone dosage . The above discussion clearly indicates that ozone dosage needs to be optimized in an ozonation process and that a number of experimental factors could influence the removal rate.
Though the ozonation process is effective for treating some organic compounds, a key problem is the accumulation of refractory compounds which interfere with the mineralization of the organic matter present in water. Some compounds were even found to be refractory to the ozonation process [15, 82, 83]. To improve its efficiency, ozonation was therefore combined with other oxidants. The combination of single oxidants can offer very effective treatment by producing more hydroxyl radicals.
It was reported that ozone in the presence of UV light enhances the decomposition rate of pollutants present in wastewater. The hydroxyl radicals generated in the UV/O3 process are shown in (2). Decomposition may proceed in three different ways: (i) by ozonation, (ii) by direct UV photolysis, and (iii) by photolysis of ozone which generates hydroxyl radicals. The detail of the ozone photolysis mechanism is shown in (6)–(10) . These combined processes should synergistically accelerate the degradation rate in the UV/O3 process compared to the ozonation process. Many authors found that the UV/O3 process is more efficient than ozonation in organic compounds degradation and mineralization [85–88]:
Recent studies also combined H2O2, and TiO2 with the UV/O3 process [89–91]. These processes are more efficient than O3/UV alone due to their synergistic effect. The dosage of H2O2 in the O3/H2O2 process needs to be optimized, however. For example, Medellin-Castillo et al. studied diethyl phthalate degradation using the O3/H2O2 process with 0.45 to 1.80 mM of H2O2, noting a linear relation with the degradation rate . Kwon et al. studied 1,4-dioxane degradation with H2O2/O3 (w/w) ratios of 0, 0.25, 0.5, 0.75, 1, 1.25, and 1.5. They found the optimum dosage ratio to be 0.5 and noted a strong retardation effect at a ratio of 1.5 . Similar results were also reported by Kusic et al. for the mineralization of phenol . This could be because an excess of hydrogen peroxide could react with the hydroxyl radical produced during the decomposition process as shown in (11) and (12). So the H2O2 dosage needs to be optimized in the degradation process. Many authors have used various combined AOPs for pollutant degradation and their conclusions are summarized in Table 2. Consider
The presence of transition metal ions such as Mn2+, Co2+, Ag+, and Fe2+ in the ozonation process has significant catalytic effects in producing hydroxyl radical [92, 93]. Abd El-Raady and Nakajima studied the degradation of formic, oxalic, and maleic acids in the presence of first row transition metal ions such as Co2+, Ni2+, Mn2+, Cu2+, Zn2+, Cr3+, and Fe2+ and compared the process efficiency with the O3 and O3/H2O2 processes . They concluded that the presence of Co2+ and Mn2+ ions has the highest catalytic activity for the decomposition of oxalic acid and that O3/Co2+ and O3/Mn2+ are more efficient than the O3/H2O2 process. Similarly, Cortes et al. reported that the O3/Mn2+ and O3/Fe2+ processes were more effective in the removal of organochloride compounds than the O3/Fe3+ and O3/high pH systems . Beltrán et al. found that the presence of Co2+ in water significantly enhances the ozonation rate of oxalic acid at acidic pH and that catalytic ozonation proceeds through the formation of a Co(HC2O4)2 complex . Heterogeneous catalytic ozonation has received increasing attention due to its potentially higher effectiveness in the degradation of recalcitrant pollutants [95–101].
3. Fenton and Photo-Fenton Based AOPs
3.1. Fenton Reaction
The Fenton process has its root in the finding reported in 1894 that ferrous ion strongly elevated the oxidation of tartaric acid by hydrogen peroxide . In the Fenton process, hydrogen peroxide is added to wastewater in the presence of ferrous salts, generating species that are strongly oxidative with respect to organic compounds. •OH is traditionally regarded as the key oxidizing species in Fenton processes. The Fenton process mechanism is quite complex and is described in detail with equations in the literature [103, 104]. In summary, the classic Fenton free radical mechanism in the absence of organic compounds mainly involves the following sequence of reactions [102, 105]:
•OH radicals are rapidly generated through (14). In the above reactions, iron cycles between Fe2+ and Fe3+ and plays the role of catalyst. The net reaction of (14)–(20) is the decomposition of H2O2 into water and O2 catalyzed by iron as follows (21):
As iron(II) acts as a catalyst, it has to be regenerated, which seems to occur through the following scheme 
Generally speaking, Fenton’s oxidation process is composed of four stages including pH adjustment, oxidation reaction, neutralization and coagulation, and precipitation. The organic substances are removed at two stages of oxidation and coagulation [106, 107]. •OH radicals are responsible for oxidation, and coagulation is ascribed to the formation of ferric hydroxo complexes [107, 108]. The relative importance of oxidation and coagulation depends primarily on the H2O2/Fe2+ ratio. Chemical coagulation predominates at a lower H2O2/Fe2+ ratio, whereas chemical oxidation is dominant at higher H2O2/Fe2+ ratios [107, 109]. Wang et al.  and Lau et al.  reported that, in Fenton treatment of biologically stabilized leachate, oxidation and coagulation were responsible for approximately 20% and 80% of overall COD, removal respectively. Fenton oxidation has been tested with a variety of synthetic wastewaters containing a diversity of target compounds, such as phenols [112–114], chlorophenols , formaldehyde , 2,4-dinitrophenol , 2,4,6-trinitrotoluene , 2,4-dinitrotoluene, chlorobenzene, tetrachloroethylene , halomethanes, amines, and hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) . Many chemicals are refractory to Fenton oxidation, however, such as acetic acid, acetone, carbon tetrachloride, methylene chloride, oxalic acid, maleic acid, malonic acid, n-paraffins, and trichloroethane . It has been demonstrated that these compounds are resistant under the usual mild operating conditions of Fenton oxidation [114, 120, 121]. In addition to these basic studies, the process has been applied to industrial wastewaters (such as chemical, pharmaceutical, textile, paper pulp, cosmetic, and cork processing wastewaters), sludge, and contaminated soils  resulting in significant reductions of toxicity, improvement of biodegradability, and colour and odour removal .
The oxidation rate was influenced by many factors such as pH value, Fe2+ : H2O2 ratio, and the amount of iron salt. Some of these parameters are discussed in detail in the following sections. The Fenton process seems to be the best compromise because it is technologically simple, there is no mass transfer limitation (homogeneous nature), and both iron and hydrogen peroxide are cheap and nontoxic. From the economic point of view, using the Fenton process as a pretreatment can lower the cost and improve biological treatment efficiency .
A batch Fenton reactor essentially consists of a pressurized stirred reactor with metering pumps for the addition of acid, a base, a ferrous sulphate catalyst solution and industrial strength (35–50%) hydrogen peroxide. It is recommended that the reactor vessel be coated with an acid resistant material, because Fenton’s reagent is very aggressive and corrosion can be a serious problem. The pH of the solution must be adjusted to maintain the stability of the catalyst, as at pH 6 iron hydroxide is usually formed. For many chemicals the ideal pH for the Fenton reaction is between 3 and 4, and the optimum catalyst to peroxide ratio is usually 1 : 5 wt/wt. Reactants are added in the following sequence: wastewater followed by dilute sulphuric acid catalyst in acidic solutions, base or acid for the adjustment of pH at a constant value, and lastly hydrogen peroxide (which must be added slowly, maintaining a steady temperature). Since wastewater compositions are highly changeable, there are some design considerations to enable the Fenton reactor to operate within flexible parameters. The discharge from the Fenton reactor is fed into a neutralizing tank to adjust the pH of the stream, followed by a flocculation tank and a solid-liquid separation tank for adjusting the TDS (total dissolved solids) content of the effluent stream. A schematic representation of the Fenton oxidation treatment is shown in Figure 2 .
As mentioned above, Fenton oxidation was applied to wastewater treatment based on the following observed optimum pH conditions, since this has been shown to affect the degradation of pollutants significantly [106, 124, 125]. The best value pH has been observed to be 2.8–3 in the majority of cases; [116, 126, 127], hence this is the recommended operating pH. At lower pH (pH = 2.5), the formation of (Fe(II) (H2O))2+ occurs, which reacts more slowly with hydrogen peroxide, producing a smaller amount of reactive hydroxyl radicals by reducing the degradation efficiency .
Furthermore, the scavenging effect of hydroxyl radicals by hydrogen ions becomes important at a very low pH, at which the reaction of Fe3+ with hydrogen peroxide is also inhibited. At an operating pH of >3, the decomposition rate decreases because of the decreased free iron species in the solution, probably due to the formation of Fe(II) complexes with the buffer inhibiting the formation of free radicals. At a pH higher than 3, Fe3+ starts precipitating as ferric oxyhydroxides and breaks down the H2O2 into O2 and H2O [124, 128], inhibiting the generation of ferrous ions. Additionally, the oxidation potential of •OH radical is known to decrease with an increase in pH .
Usually the rate of degradation increases with an increased concentration of ferrous ions , though the increase is sometimes observed to be marginal above a certain concentration [106, 129]. Additionally, an enormous increase in ferrous ions will lead to an increased unutilized quantity of iron salts, contributing to increased TDS content in the effluent treatment, which is not permitted. Thus laboratory scale studies are required to establish the optimum loading of ferrous ions under similar conditions, unless data are available in the open access literature .
The concentration of hydrogen peroxide plays a more crucial role in the overall efficacy of the degradation process. Usually it has been observed that the percentage degradation of the pollutant increases with an increased dosage of hydrogen peroxide [106, 129]. Care should be taken however in selecting the operating oxidant dosage. The residual hydrogen peroxide contributes to COD, so an excess amount is not recommended. The presence of hydrogen peroxide is also harmful to many microorganisms and affects the overall degradation efficiency significantly where Fenton oxidation is used as a pretreatment to biological oxidation. One more negative effect of hydrogen peroxide, if present in large quantities, is that it acts as a scavenger for the generated hydroxyl radicals. Thus hydrogen peroxide loading should be adjusted so that the entire amount is utilized. This can be decided based on laboratory scale studies with the effluent in question .
It should be noted that the dose of H2O2 and the concentration of Fe2+ are two relevant and closely related factors affecting the Fenton process. The H2O2 dose has to be fixed according to the initial pollutant concentration. An amount of H2O2 corresponding to the theoretical stoichiometric H2O2 to chemical oxygen demand (COD) ratio is frequently used , although it depends on the response of the specific contaminants to oxidation and on the objective pursued in terms of reducing the contaminant load. Usually a lower initial pollutant concentration is favoured , but the negative effects of treating a large quantity of effluent need to be analyzed before the dilution ratio can be set. For real industrial wastes, some dilution is often essential before any degradation is observed using Fenton oxidation .
As noted above, as the maximum degradation rates are observed at a pH of approximately 3, the operating pH should be maintained constant around this optimum value. The type of buffer solution used also affects the degradation process . Acetic acid/acetate buffer provides maximum oxidation efficiency, at least as observed for phosphate and sulphate buffers. This can be attributed to the formation of stable Fe3+ complexes under these conditions .
Not many studies are available depicting the effect of temperature on degradation rates and ambient conditions can safely be used with good efficiency . Besides, reaction temperature is another crucial parameter in the Fenton process. In principle, increasing the temperature should enhance the kinetics of the process, but it also favours the decomposition of H2O2 towards O2 and H2O. This increases at a rate of around 2.2 times each 10°C in the range of 20–100°C . Oxidation with Fenton’s reagent has already been proved effective and promising for the destruction of several compounds and consequently for the treatment of a wide range of wastewaters, as described in several reviews (e.g. [102, 104, 116, 123, 131, 132]). Table 3 summarizes recent Fenton processes for some wastewater treatments.
3.2. Photo-Fenton Processes
The photo-Fenton process, as its name suggests, is rather similar to the Fenton one but also employs radiation [102, 104, 123, 133]. The photo-Fenton reaction is also well known in the literature [104, 134] as an efficient and inexpensive method of wastewater and soil treatment [104, 135]. Photo-Fenton process is known to be capable of improving the efficiency of dark Fenton or Fenton-like reagents by means of the interaction of radiation (UV or Vis) with Fenton’s reagent . This technique has been suggested as feasible and promising for removing pollutants from natural and industrial waters and increasing the biodegradability of chlorophenols when used as a pretreatment method to decrease water toxicity . Some of its most innovative applications include oxalate as a ligand of iron ions [104, 137].
The effectiveness of photo-Fenton processes is attributed to the photolysis of Fe(III) cations in acidic media yielding Fe(II) cations (24), in conjunction with reaction between Fe(II) and •OH to yield hydroxyl radicals (Fenton’s reaction, step 24):
In this process, the photolytic decomposition of Fe(OH)2+ (24) is accelerated, providing an additional source of highly oxidative hydroxyl radicals compared to the “simple” Fenton process . The photo-Fenton process produces more hydroxyl radicals than the conventional Fenton method (Fe(II) with hydrogen peroxide) or photolysis, thus promoting organic pollutant degradation rates. This process consists of a combination of Fenton reagents (Fe2+/H2O2) and light energy [138, 139] and thus of two reactions :
The first reaction is a reaction of Fe2+ with H2O2, which generates the powerful reactive species •OH radicals and oxidizes Fe2+ to Fe3+. In other words, the hydroxyl radical generation in Fenton processes is due to the iron catalyzed decomposition of H2O2. The first photo-Fenton reaction causes the formation of hydroxyl radicals. The second reaction of the photo-Fenton process is a reaction of Fe3+ with water, which occurs when light is used at a wavelength from 300 nm to 650 nm. This generates •OH radicals and reduces Fe3+ to Fe2+. These two oxidation-reduction reactions occur repeatedly and completely mineralize organic pollutants to CO2 and H2O :
The oxidation power of the photo-Fenton process is attributed to the generation of OH radicals. Without irradiation, a Fenton-like reaction occurred instead of a photo-Fenton reaction. The Fenton-like reaction is a reaction of Fe3+ with H2O2, which causes the reduction of Fe3+ to Fe2+:
Since Reaction (30) occurs instead of Reaction (27), organic pollutants are mineralized even without irradiation. It should be noted, however, that Reaction (30) is rather slower than Reaction (27). Thus the degradation rate under dark conditions is rather lower than that of the photo-Fenton reaction . Figure 3 shows the reaction pathways for the process starting with the primary photoreduction of the dissolved Fe(III) complexes to Fe(II) ions followed by Fenton’s reaction and the subsequent oxidation of organic compounds. Additional hydroxyl radicals generated in the first step also take part in the oxidation reaction .
Appropriate implementation of the photo-Fenton treatment depends mainly on the operating variables—H2O2/COD molar ratio, H2O2/Fe2+ molar ratio, and irradiation time. The conventional method is to optimize the operating variables by changing one factor at a time; that is, a single factor is varied while all other factors are kept unchanged for a particular set of experiments. Likewise, other variables are individually optimized through single-dimensional searches, which are time consuming and incapable of reaching the actual optimum as interaction among variables is not taken into consideration . Some illustrative works from recent years are discussed in detail in Table 4.
4. UV/H2O2 Process
Like other AOPs, the oxidizing ability of UV/H2O2 may be attributed to the formation of •OH, , and as reflected by their mechanistic pathways (Reactions (1), (19)–(23)). In fact, the AOP occurs via a reaction with •OH radicals, produced by UV irradiation of H2O2. The molar absorptivity of hydrogen peroxide is low at 253.7 nm, about 20 M−1 cm−1, and •OH radicals are formed per incident photon absorbed . At this wavelength, the photolysis rate of aqueous hydrogen peroxide is about 50 times slower than that of ozone . This technique requires a relatively high dose of H2O2 and/or a much longer UV exposure time than, for example, the UV/O3 process. In contrast, the rate of photolysis of hydrogen peroxide has been found to be pH dependent and increases when more alkaline conditions are used, because at 253.7 nm peroxide anions may be formed, which display a higher molar absorptivity than hydrogen peroxide, namely, 240 M−1 cm−1 [104, 142]. In this AOP, the formation of •OH radicals is directly facilitated by the photolysis of H2O2 . The radicals, which are formed by the homolytic splitting of the oxidant’s O–O bonds, transform the chemical structures of target chelating agents [143, 144]. Consider the following: Initiation (Rate Constant) Propagation  (Rate Constant) Termination  (Rate Constant) It is important to note that the effectiveness of UV/H2O2 systems depends on various conditions that affect their ability to degrade chelating agents. The variables include the type and concentration of contaminants or dissolved inorganic substances (such as carbonates and iron cations), organic substances present in surface water, light transmittance in solutions (as indicated by turbidity or colour), pH, temperature, and the optimum oxidant dose . An excessive concentration of H2O2 would act as a radical scavenger, slowing down the rate of oxidation , while a low concentration of H2O2 insufficiently forms OH radicals in aqueous solutions, leading to a slower oxidation rate [143, 147]. The UV/H2O2 process is sensitive to the scavenging effects of carbonate ions at a pH ranging from 8 to 9. Furthermore, the UV/H2O2 process requires a long UV exposure time with a powerful output at a wide range of wavelengths. Nevertheless this treatment is more economically attractive than the UV/O3 process, due to its lower energy consumption [143, 148].
Tubular reactor configurations are usually employed for direct photolysis and photo-Fenton processes or processes based on H2O2/UV reagent, in order to achieve a good interaction between CPs, other intermediates, and radiation [104, 149]. Also various lamps are employed to generate the radiation supplied to CP samples for direct UV photolysis and for techniques based on UV/H2O2, UV/O3, photo-Fenton processes, and photocatalysis. The various commercial radiation sources employed include high, medium, and low pressure mercury vapour lamps for the generation of UV radiation [149–151] and solar-simulated xenon lamps as a source of visible radiation . The lamp can be located either in an axial position housed by a sleeve  or vertically, in its centre . The typical findings observed in the UV/H2O2 process are listed in Table 5.
Ultimate oxidation of CPs to carbon dioxide and water has rarely been obtained under typical test conditions. As summarized in Table 4, typical half-life times are between 0.3 and 20.1 minutes for CP degradation, depending on the initial concentration of CP and hydrogen peroxide, the intensity of radiation, and the degree of chlorination. It is observed that the degradation rates increase when the number of chlorine substituents decreases .
Recent developments in various homogeneous AOPs have been analysed comprehensively. The principle of individual and combined AOPs and their efficiency on the degradation of various pollutants was discussed. The influence of various experimental parameters such as oxidant dosage, solution pH, flow rates, substrate concentrations, water matrix, and light intensity on the AOPs was explored. This review also listed various AOPs applied for the degradation of contaminants under different experimental conditions. Combined AOPs substantially enhanced the degradation rate by generating more reactive radicals under suitable conditions. The optimum oxidant dosage and solution for efficient removal were reported.
Conflict of Interests
The authors declare that there is no conflict of interests regarding the publication of this paper.
The authors wish to thank the National Science Council (NSC) in Taiwan for their financial support under the Contract no. NSC-101-2221-035-031-MY3. The Laboratory of Green Chemistry, Mikkeli, Finland, and Water and Environmental Technology (WET) Center, Temple University, are also gratefully acknowledged for their partial financial support of this study.
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