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Heterogeneous photocatalytic degradation of phenols in wastewater A review on current status a


Desalination 261 (2010) 3–18

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Desalination
j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / d e s a l

Heterogeneous photocatalytic degradation of phenols in wastewater: A review on current status and developments
Saber Ahmed a,?, M.G. Rasul a, Wayde N. Martens b, R. Brown c, M.A. Hashib d
a

Faculty of Science, Engineering and Health, CQ University, QLD 4702, Australia Queensland University of Technology, Discipline of Chemistry, Faculty of Science and Technology, Australia Queensland University of Technology, School of Engineering Systems, Australia d Department of Ecological Engineering, Toyohashi University of Technology, Japan
b c

a r t i c l e

i n f o

a b s t r a c t
In recent years, there has been an enormous amount of research and development in the area of heterogeneous photocatalytic water puri?cation process due to its effectiveness in degrading and mineralising the recalcitrant organic compounds as well as the possibility of utilising the solar UV and visible spectrum. One hundred and twenty recently published papers are reviewed and summarised here with the focus being on the photocatalytic oxidation of phenols and their derivatives, predominant in waste water ef?uent. In this review, the effects of various operating parameters on the photocatalytic degradation of phenols and substituted phenols are presented. Recent ?ndings suggested that different parameters, such as type of photocatalyst and composition, light intensity, initial substrate concentration, amount of catalyst, pH of the reaction medium, ionic components in water, solvent types, oxidising agents/electron acceptors, mode of catalyst application, and calcination temperatures can play an important role on the photocatalytic degradation of phenolic compounds in wastewater. Extensive research has focused on the enhancement of photocatalysis by modi?cation of TiO2 employing metal, non-metal and ion doping. Recent developments in TiO2 photocatalysis for the degradation of various phenols and substituted phenols are also reviewed. ? 2010 Elsevier B.V. All rights reserved.

Article history: Received 18 December 2009 Received in revised form 30 March 2010 Accepted 29 April 2010 Available online 9 June 2010 Keywords: Phenols Substituted phenols Photocatalysis Water puri?cation

Contents 1. 2. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Photocatalytic oxidation process . . . . . . . . . . . . . . . . . . . 2.1. Principle of photocatalysis . . . . . . . . . . . . . . . . . . . 2.2. Phenol, chlorophenol and nitrophenol in photocatalytic medium . 3. In?uence of parameters on the photocatalytic degradation of phenols . . 3.1. Comparison of catalyst and substituted group . . . . . . . . . . 3.2. Light intensity and wavelength . . . . . . . . . . . . . . . . . 3.3. Initial concentration . . . . . . . . . . . . . . . . . . . . . . 3.4. Catalyst loading . . . . . . . . . . . . . . . . . . . . . . . . 3.5. Medium pH . . . . . . . . . . . . . . . . . . . . . . . . . . 3.6. Co-occurring substances . . . . . . . . . . . . . . . . . . . . 3.7. Oxidants/electron acceptor . . . . . . . . . . . . . . . . . . . 3.8. Calcination temperature . . . . . . . . . . . . . . . . . . . . 3.9. Doping and mixed semiconductor . . . . . . . . . . . . . . . 4. Conclusion. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgement . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4 3 4 4 5 5 5 7 8 8 9 10 11 12 14 15 16 16

? Corresponding author. Fax: +61 7 4930 9382. E-mail address: s.ahmed@cqu.edu.au (S. Ahmed). 0011-9164/$ – see front matter ? 2010 Elsevier B.V. All rights reserved. doi:10.1016/j.desal.2010.04.062

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1. Introduction Recycling of wastewater ef?uent is recognised to be a strategic approach in a sustainable water management portfolio by water utilities both in Australia and in other parts of the world to minimise the growing water demand in a water-scarce environment [1–3]. The widespread occurrence of phenols in waste water and associated environmental hazards has heightened concern over public health [4]. Phenols and their derivatives are well known for their biorecalcitrant and acute toxicity. Phenols are being introduced continuously into the aquatic environment through various anthropogenic inputs. The presence of toxic organic compounds in storm and waste water ef?uent is reported to be a major impediment to the widespread acceptance of water recycling [5,6]. Further, their variety, toxicity and persistence can directly impact the health of ecosystems and present a threat to humans through contamination of drinking water supplies e.g., surface and ground water [3,4]. In response it has become a challenge to achieve the effective removal of persistent organic pollutants from waste water ef?uent to minimise the risk of pollution problems from such toxic chemicals and to enable its reuse. Consequently, considerable efforts have been devoted to developing a suitable puri?cation method that can easily destroy these bio-recalcitrant organic contaminants. Due to their incomplete removal during wastewater treatment, they are ubiquitous in secondary wastewater ef?uents, rivers and lakes at low concentration. Despite their low concentration, these contaminants are a major health concern because of their extremely high endocrine disrupting potency and genotoxicity [7]. These ?ndings enunciate the necessity for further research on the removal of trace contaminants to minimise their accumulation, particularly prior to indirect or direct reuse of reclaimed water. Conventional wastewater puri?cation systems such as activated carbon adsorption, membrane ?ltration, chemical coagulation, ion exchange on synthetic adsorbent resins, etc., also generate wastes during the treatment of contaminated water, which requires additional steps and cost. In recent years, the heterogeneous photocatalytic oxidation (HPO) process employing TiO2 and UV light has emerged as a promising new route for the degradation of persistent organic pollutants, and produces more biologically degradable and less toxic substances [8,9]. This process is largely dependent on the in-situ generation of hydroxyl radicals under ambient conditions which are capable of converting a wide spectrum of toxic organic compounds including the non-biodegradable ones into relatively innocuous end products such as CO2 and H2O. In the HPO process, destruction of recalcitrant organics is governed by the combined actions of a semiconductor photocatalyst, an energetic radiation source and an oxidizing agent. Moreover, the process can be driven by solar UV or visible light. Near the earth's surface, the sun produces 0.2–0.3 mol photons m? 2 h? 1 in the range of 300–400 nm with a typical UV ?ux of 20–30 W m2. This suggests using sunlight as an economically and ecologically sensible light source [10]. As a result, development of an ef?cient photocatalytic water puri?cation process for large scale applications has received considerable attention. Although heterogeneous photocatalysis appears in numerous forms, the photocatalytic degradation of phenol and substituted phenols in wastewater has recently been most widely investigated. Several reviews [11,12] focusing brie?y on the available technologies for the abatement of phenolic compounds in wastewater have recently been published. In light of the basic and applied research reviewed, the photocatalytic oxidation method appears to be a promising route for the treatment of wastewater contaminated with phenol and phenolic compounds. However, the major drawbacks still encountered are low quantum ef?ciency due to inef?cient visible light harvesting [13], the design of the photoreactor [14], the recovery and reuse of titanium dioxide [15], the generation of toxic intermediates [16,17], and concern about

catalyst deactivation [18]. Information from various investigations suggests that the photocatalytic degradation of phenol and its derivatives is mainly dependent on the solution pH, catalysts and their composition, organic substrate type and concentration, light intensity, catalyst loading, ionic composition of waste water, types of solvent, oxidant concentration, and calcination temperatures. Understanding the impacts of various parameters on the photocatalytic degradation ef?ciency is of paramount importance from the design and operational points of view when choosing a sustainable, ef?cient technique for the treatment of waste water. This paper aims to review and summarise the role of important operating parameters on the photocatalytic degradation of phenols in wastewater together with recent achievements. Recent research on the modi?ed TiO2 photocatalysis under visible and/or solar light aimed at improving phenols degradation in wastewater by means of metal, non-metal and ion doping is also highlighted in this review. The existing limitation and future research needs associated with the treatment technology pertaining to the contaminant of interest are also discussed.

2. Photocatalytic oxidation process 2.1. Principle of photocatalysis In the photocatalytic oxidation process, organic pollutants are destroyed in the presence of semiconductor photocatalysts (e.g., TiO2 and ZnO), an energetic light source, and an oxidising agent such as oxygen or air. Fig. 1 illustrates the mechanism of TiO2 photocatalysis. Only photons with energies greater than the band-gap energy (ΔE) can result in the excitation of valence band (VB) electrons which then promote the possible reactions. The absorption of photons with energy lower than ΔE or longer wavelengths usually causes energy dissipation in the forms of heat. The illumination of the photocatalytic surface with suf?cient energy leads to the formation of a positive hole (h+) in the valence band and an electron (e?) in the conduction band v (CB). The positive hole oxidizes either pollutant directly or water to produce ·OH radicals, whereas the electron in the conduction band reduces the oxygen adsorbed on the catalyst (TiO2).

Fig. 1. Schematic diagram illustrating the principle of TiO2 photocatalysis.

S. Ahmed et al. / Desalination 261 (2010) 3–18

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The activation of TiO2 by UV light can be represented by the following steps: TiO2 ? hν →e ? h e ? O2 →O2 ?
? ? ? ?

?2:1? ?2:2?

In this reaction, h+ and e? are powerful oxidizing and reductive agents respectively. The oxidative and reductive reaction steps are expressed as: Oxidative reaction: h ? Organic→CO2 h ? H2 O→ OH ? H
? ? ? ?

?2:3? ?2:4?

Reductive reaction:
?

OH ? Organic→CO2

?2:5?

Hydroxyl radical generation by the photocatalytic oxidation process is shown in the above steps. In the degradation of organic pollutants, the hydroxyl radical, which is generated from the oxidation of adsorbed water where it is adsorbed as OH?, is the primary oxidant; and the presence of oxygen prevents the recombination of an electron–hole pair. In the photocatalytic degradation of pollutants, when the reduction process of oxygen and the oxidation of pollutants do not advance simultaneously, there is an electron accumulation in the CB, thereby causing an increase in the rate of recombination of e? and h+. Thus it is of paramount importance to prevent electron accumulation in ef?cient photocatalytic oxidation. In photocatalysis, TiO2 has been studied extensively because of its high activity, desirable physical and chemical properties, low cost, and availability. Of three common TiO2 crystalline forms, anatase and rutile forms have been investigated extensively as photocatalysts. Anatase has been reported to be more active as a photocatalyst than rutile. Different light sources such as UV lamps and solar radiation have been used in previous investigations into the photocatalysis of various pesticide and herbicide derivatives dominant in storm water and wastewater ef?uent.

converted to acetylene, maleic acid, carbon monoxide and carbon dioxide [20]. Chlorophenols are water pollutants of moderately toxicity and are suspected of carcinogenic properties. They originate from the natural degradation of chlorinated herbicides, the chlorination of phenolic substances and disinfection of water. Nitrophenols are water pollutants of high toxicity and they are released into the water environment due to the synthesis and use of organo-phosphorous pesticides, azo dyes, and some medical goods. The main byproducts detected during their photocatalytic degradation are 4-nitrocatechol, benzoquinone, hydroquinone and a number of organic acids. Fig. 2 illustrates the photocatalytic degradation of phenol and the formation of intermediates. The ?OH radical attacks the phenyl ring of the phenol yielding catechol, resorcinol, 1,2,3-benzenetriol and hydroquinone, then the phenyl rings in these compounds disintegrate to give malonic acid, then short-chain organic acids such as, maleic, oxalic, acetic, formic 3-hydroxy propyl carboxylic acid, 2-hydroxy propanal, 2-hydroxy-ethanoic acid glycol acid, ?nally CO2 and H2O. H? produced during the attack of bonds by ?OH is reported to be an important active free radical in the degradation process. During the process, H+ or H? is scavenged by oxygen to form HO2? radicals, which eventually convert to ?OH radicals. Therefore, the principal reaction leading to organics decomposition would be the one with ?OH radicals. 3. In?uence of parameters on the photocatalytic degradation of phenols 3.1. Comparison of catalyst and substituted group The photocatalytic activity of TiO2 is dependent on surface and structural properties of the semiconductor such as crystal composition, surface area, particle size distribution, porosity, band gap and surface hydroxyl density. Particle size is of primary importance in heterogeneous catalysis, because it is directly related to the ef?ciency of a catalyst through the de?nition of its speci?c surface area. A number of commercially available catalysts have been investigated for the photocatalytic degradation of phenolic compounds and dyes in aqueous environments. Table 2 shows the speci?cation and characteristics of some commercial TiO2 samples. The photocatalyst titanium dioxide Degussa P-25 has been widely used in most of the experimental conditions; other catalyst powders, namely, Hombikat UV100, PC 500 and TTP were also used for degradation of toxic organic compounds. P-25 contains 75% anatase and 25% rutile with a speci?c BET surface area of 50m2/g and a primary particle size of 20 nm [25]. Hombikat UV 100 consists of 100% pure and smaller anatase with a speci?c BET surface area of 250 m2/g and a primary particle size of 5 nm [25]. The photocatalyst PC 500 has a BET surface area of 287 m2/ g with 100% anatase and a primary particle size of 5–10 nm [25], and TiO2 obtained from TTP, India has a BET surface area of 9.82 m2/g. It has been demonstrated that the degradation rate of dyes proceeds much more rapidly in the presence of Degussa P-25 as compared to other photocatalysts. The ef?ciency of photocatalysts was shown to follow the order: P25 N UV100 N PC500 N TTP for the degradation of various dyes [26–32]. The photocatalytic activity of Degussa P25 was reported to be higher due to slow recombination between electrons and holes whereas Hombikat UV 100 is reported to have a high photoreactivity due to fast interfacial electron transfer rate. The differences in the photocatalytic activity are related to the variations in the BET surface, impurities, existence of structural defects in the crystalline framework or density of hydroxyl groups on the catalyst surface. These factors could in?uence the adsorption behaviour of a pollutant or degradation intermediates and the life time and recombination rate of electron–hole pairs [27]. Some of the illustrative works in recent years is depicted in Table 3. Priya and Madras [33] compared the photocatalytic degradation of nitrophenols using

2.2. Phenol, chlorophenol and nitrophenol in photocatalytic medium Phenol along with its chlorophenol and nitrophenol derivatives is highly soluble in water. The solubility of phenol, o-chlorophenol, m-chlorophenol, p-chlorophenol, o-nitrophenol, m-nitrophenol and p-nitrophenol is shown in Table 1. The main reaction site for their destruction during photocatalysis is the bulk liquid, where the attack of hydroxyl radicals on the ring carbons results in various oxidation intermediates. Hydroquinone, catechol, and p-benzoquinone are reported to be the major intermediates formed during the photocatalytic degradation of phenol [19]. Reaction intermediates such as chloro-hydroquinone, 4-chlorocatechol and resorcinol are eventually

Table 1 Solubility of phenol and its chloro and nitro-derivatives in water. Compound Phenol o-chlorophenol m-chlorophenol p-chlorophenol o-nitrophenol m-nitrophenol p-nitrophenol Solubility (g/100 g) 9.48 2.04 2.25 2.77 0.21 2.19 1.34 Ref. [21] [22] [22] [22] [23] [23] [23]

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Fig. 2. Photocatalytic degradation of phenol by nanomaterial TiO2 in wastewater [24].

combustion synthesised nano-TiO2 and Degussa P-25. The photodegradation kinetics was ?rst order. The photocatalytic degradation rates were considerably higher in combustion synthesized TiO2 compared to that of P-25. For both catalysts, the degradation rate was shown to follow the order 4-Nitrophenol N 2-Nitrophenol N 3-Nitrophenol N 2,4-Ditrophenol. The position of substitution is reported to affect the rate of degradation. Shukla et al. [34] compared the photocatalytic degradation ef?ciency of phenol, dichlorophenol (DP) and trichlorophenol (TCP) at 0.266 mM solution concentration in the presence of UV–vis/ZnO/persulfate. Degradation of the tested compounds under simulated solar light alone was demonstrated to be negligible due to poor absorbance of light in the available wavelength. In comparison to phenol, chlorophenols degradation was signi?cantly lower under the conditions examined. The order of the degradation rate was found to be as phenol N DCP N TCP. The photocatalytic activity of ZnO was observed to be higher than TiO2 for the degradation of phenolic contaminants under UV–vis light. Salah et al. [35] compared the ef?ciency of three commercial catalysts Degussa P25 (20 nm), TiO2-A1 (160 nm) and TiO2-A2 (330 nm) available in anatase form, and ZnO for the degradation of phenol. After 5 h reactions, the order of the ef?ciency was shown to follow ZnO N P25 N TiO2-A1 N TiO2-A2. The observed variation in the ef?ciency is related to the structure, diameter of particles and electrical properties of the photocatalyst. Peiro et al. [36] observed that the photocatalytic degradation of phenol and ortho-substituted phenolic compounds were shown to follow the order: Guaiacol N 2-chlorophenol = Phenol N Catechol. Kusvuran et al. [37] studied the photocatalytic degradation of 2,4,6-trimethylphenol (TMP), 2,4,6-trichlorophenol (TCP), 2,4,6tribromophenol (TBP), 2,4-dimethylphenol (DMP), 2,4-dichlorophenol (DCP) and 2,4-dibromophenol (DBP). Under the conditions tested, degradation of these tri-substituted phenols was found to be in the following order: TCP N TBP N TMP and the degradation of di-substituted phenols was shown to be in the order: DCP N DBPN DMP. The degradation rate of the tested compounds was observed to be strongly dependent

on the adsorption over catalyst, the electronic nature and number of the substituents and their position in the aromatic ring. Lachheb et al. [38] reported that in comparison to PC 500, P-25 was more ef?cient for the degradation of phenols and poly nitrophenols (4-NP, 2,4-DNP and 2,4,6-TNP) in the presence of either arti?cial or solar light. The degradation followed ?rst order kinetics. The photocatalytic degradations of the tested compounds were shown to be in the following order: 2,4,6-TNPN 2,4-DNPN 4-NPN Phenol. However, for PC 500 supported on Ahlstrom paper 1048, the order is the opposite — PhenolN
Table 3 Some representative work on the photocatalytic degradation of organic contaminants. Pollutant Phenol Experimental conditions Degradation rate 92% degradation achieved in 6 h Ref. [35]

C0 phenol = 50ppm, TiO2 = 2g/l, under UV irradiation (λmax = 366 nm) Tri-substituted TiO2 = 0.5 g/l in 200 ml, phenol; TCP, TBP, TMP Co = 0.?0.5 mM

Phenol, DNP, 4-NP, TNP

Phenol, DCP, TCP

Co = 30 mg/l, I = 125W MP Hg, λ N 290 nm, TiO2 = 0.5 g/l, C0Phenol = 25ppm, C0DCP = 44 ppm C0TCP = 52 ppm, ZnO = 0.4 g/l, Persulfate = 2 g/l

Chrysoidine R 1 , Acid C0 = 0.22–0.25 mM, Red 292 125 W Mp Hg, TiO2 = 1 g/l, pH= 4.3– 5.2

Amaranth1, Bismarck2

Table 2 Speci?cation and characteristics of commercial TiO2 samples. TiO2 sample Degussa P-25 Millennium PC 500 Hombikat UV 100 TTP N/A = Not available. BET surface area Particle size (nm) Composition (m2/g) 50 287 250 9.82 21 5–10 5 N/A 75% Anatase, 25% Rutile Anatase Anatase N/A

C0 = 0.4–0.5 mM, 125 W Mp Hg, TiO2 = 1 g/l, pH = 3.25–5.68 Acid blue 451, Xelenol C0 = 0.3–0.4 mM, orange2 125 W Mp Hg, TiO2 = 1 g/l, pH = not available Bromothymol blue C0 = 0.25 mM, TiO2 = 1 g/l, 125W Mp Hg, pH = not available Acridine Orange1, C0 = 1.0 mM, 2 Ethidium bromide TiO2 = 1 g/l, 125 W Mp Hg, pH = not available

100―56% TCP removal obtained in 1 h. 70―44% TBP removal achieved in 1 h. 60―22% TMP removal in 1 h. Complete degradation of all nitro phenols occurs between 2 and 7 h 85% degradation of TNP occurs in 150 min. 90% degradation of DNP occurs in 150 min and 93% degradation of phenol occurs in 150 min. 92% degradation and 77% mineralisation of dye1 takes place after 1 h irradiation. 98% degradation and 88% mineralisation of dye2 occurs in 2 h Nearly complete degradation occurs in 75 min for dye1 and 40 min for dye2. About 51 and 73% degradation of dye1 and dye2 takes place in 80 min irradiation time. Nearly complete TOC removal achieved in 90 min. Complete degradation of dye1 and dye2 takes place in 75 and 195 min respectively.

[37]

[38]

[34]

[26]

[26]

[30]

[32]

[31]

S. Ahmed et al. / Desalination 261 (2010) 3–18

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4-NP N 2,4-DNPN 2,4,6-TNP. The difference in poly nitrophenols disappearance rates was related to the variation in adsorption behaviour. For both supported Degussa P25 and Millennium PC-500 photocatalysts, the maximum quantities of adsorbed phenolic compounds increase in the order: Phenolb 4-NP b DNPb TNP. Gulliard et al. [39] compared the photocatalytic ef?ciency of different catalysts with various surface areas, crystallite and particle sizes and their chemical of surface for the degradation of 4-chlorophenol. The observed order of ef?ciency based on the rate of degradation upon solar exposure was shown to be: TiONA PC 10 (9 m2/g)N P-25 (50 m2/g)N TiL COM HC 120 (120 m2/g)N Hombikat UV 100 (250 m2/g). The TiONA PC 10 photocatalyst has an activity about two times higher than Hombikat UV100 and TiL COM HC 120, and 1.4 times higher than that of P-25. All these catalysts have mainly the anatase structure. In the presence of P-25, Sevlam et al. [40] indicated that complete degradation of 4-?urophenol using P25 (BET=55 m2/g, particle size=30 nm) was achieved in 60 min while complete degradation occurred in 90 min in the presence of ZnO (BET=10 m2/g, particle size=0.1–4 ?m). The higher ef?ciency of P-25 is due to its larger surface area compared to the ZnO. The oxidizing power of hydroxyl radicals produced by these catalysts is strong enough to break C–C, C–F and C=C bonds of 4-FP adsorbed on their surfaces leading to the formation of CO2 and mineral acids. Sobana and Swaminathan [41] compared the ef?ciencies of various catalysts (ZnO, TiO2 anatase, ZnS, SnO2, Fe2O3 and CdS for the degradation of Acid red 18 under UV irradiation. The highest degradation was achieved with ZnO over TiO2 anatase catalyst due to the higher surface area of ZnO (10 m2/g) over TiO2 anatase (8.9 m2/g). SnO2, Fe2O3, CdS and ZnS have negligible activity on Acid Red 18 decolourisation due to their smaller band gaps which permit rapid recombination of hole and electron. Talebian and Nilforoushan [42] compared the photocatalytic degradation ef?ciency of methylene blue (MB) under UV irradiation. The observed order of photocatalytic ef?ciency was shown to be SnO2 b ZnO b TiO2 b In2O3. The same trend was also observed at different pH values for the photocatalytic activities of different catalysts due to their various micro structures. Zhang et al. [43] compared the ef?ciency of coupled oxide ZnO–SnO2 with ZnO or SnO2 for the degradation of methylene orange (MO). The photocatalytic ef?ciency of ZnO–SnO2 was signi?cantly higher compared to ZnO or SnO2 alone. Under the experimental conditions, the degradation of MO follows the ?rstorder reactions during the ?rst 15 min of irradiation, and the apparent rate constants for ZnO–SnO2, Degussa P25 TiO2, ZnO– SnO2/NaOH, ZnO and SnO2 are 0.0865, 0.0822, 0.0339, 0.0335 and 0.0124 min? 1, respectively. The higher activity of ZnO–SnO2 is attributed to its smaller size and the charge separation in coupled oxide. Cun et al. [44] compared the ef?cacy of ZS, pure ZnO and pure SnO2 for the photocatalytic degradation of MO. The degradation ef?ciency was reported to follow the order of ZS N ZnO N SnO2. The photocatalytic reaction by Z2S is 40.2 and 66.1% faster than those by ZS and ZnO, respectively. The higher activity of ZS was related to the larger BET of ZS than ZnO, and the suppression of electron–hole pairs' recombination. Ksibi et al. [45] studied the photocatalytic degradation of six substituted phenols (2,4,6-TNP, Hydroquinone, 2,4-DNP, Resorcinol, phenol and 4-NP) over UV irradiated TiO2. The adsorption constants were shown to decrease in the order: 2,4DNP N Hydro N 4-NP N phenol N resorcinol N 2,4,6-TNP. The higher value of 2,4-DNP (31 L/mol) was attributed to the ortho-position of nitro (–NO2) group in the aromatic ring. In contrast the lowest adsorption constant of 2,4,6-TNP (1.45 L/mol) is related to the big geometric hindrance due to the large number of nitro groups on the aromatic ring. The in?uence of the substituents on the photocatalytic degradation of phenols was reported to be Resorcinol N phenol.4-NP N 2,4-DNP N Hydro N 2,4,6-TNP. The variation in photoreactivity of these compounds is ascribed to the electronic character (electro-withdrawing or electro-donor) and to the position of the substituent groups in the aromatic ring.

3.2. Light intensity and wavelength The light intensity determines the extent of light absorption by the semiconductor catalyst at a given wavelength. The rate of initiation for photocatalysis electron–hole formation in the photochemical reaction is strongly dependent on the light intensity [46]. Light intensity distribution within the reactor invariably determines the overall pollutant conversion and degradation ef?ciency [47]. Consequently the dependency of pollutant degradation rate on the light intensity has been studied in numerous investigations for various organic pollutants [48,49]. While in some cases the rate of reaction exhibited a square root dependency on the light intensity, others observed a linear relationship between the two variables [48–50]. Ollis et al. [49] reviewed the effect of light intensity on the organic pollutant degradation rate. It has been reported [48–50] that the rate is proportional to the radiant ?ux Ф for Ф b 25 mW/cm2, and above 25 mW/cm2 the rate has been shown to be varied as Ф1/2, indicating a too high value of the ?ux and an increase of the electron–hole recombination rate. At high intensity, the reaction rate was independent of light intensity. This is likely because, at low intensity, reactions involving electron–hole formation are predominant and electron– hole recombination is insigni?cant. Puma and Yue [51] examined the effect of light wavelength on the photocatalytic degradation of 2-Chlorophenol (2-CP) using UV-A alone and individually with simultaneous UV-A, B, and-C radiation. The rate of degradation and mineralization of 2-CP were signi?cantly improved with UV-ABC radiation compared to UV-A radiation. The initial rate of 2-CP degradation was found to be 1.8 times faster under UV-ABC radiation and the degradation of 2-CP was 98% in 20 min. The enhanced degradation rate with UV-ABC was attributed to the photon ?ux with UV-ABC, which was reported to be 1.56 times higher compared to UV-A radiation. The rate improvement was related to the combined use of photolysis, photocatalysis, and synergistic effects due to associated photolysis and photocatalysis. Kaneco et al. [52] investigated the effect of light intensity on the solar photocatalytic degradation of bisphenol A in water with TiO2 on sunny and cloudy days. The degradation ef?ciency is shown to increase rapidly with increase in the light intensity up to 0.35 mW/cm2, and then the ef?ciency increased gradually. Venkatachalam et al. [53] compared the mineralisation ef?ciency of 4-CP using lamps of wavelength 365 nm and 254 nm over TiO2. The mineralisation rate at 365 nm is reported to be slightly higher than at 254 nm. Since the band gap excitation of electrons in TiO2 with 254 nm can promote electrons to the conduction band with high kinetic energy, they can reach the solid–liquid interface easily, suppressing the electron–hole recombination in comparison to 365 nm. Chiou et al. [54] studied the effect of UV light intensity (20–400W) on the phenol degradation. All the reactions followed the ?rst order kinetics. In the UV/TiO2 system, the degradation rate constants with a light intensity of 20, 100, and 400 W are 8.3 × 10? 3, 0.012 and 0.031 min? 1 respectively. Under the conditions tested, an acceptably good linear correlation exists between the apparent ?rst order rate constant and light intensity. Increasing light intensity enhances the formation of hydroxyl radical thereby improving the degradation rate. Shukla et al. [34] studied the effect of light intensity (160–330 W) on the photocatalytic degradation of phenol using ZnO and arti?cial solar light. Under the conditions tested (C0 = 25 ppm, ZnO = 0.4 g/l, Persulfate = 2 g/l), the rate of degradation is shown to be favoured by the increase in lamp power with 95% conversion being obtained within 3 h using the lamp power of 330 W as opposed to 4 h while using the lamp power of 160 W. Han et al. [55] studied the photocatalytic degradation and mineralisation of 4-CP, hydroquinone and 4-nitrophenol (4-NP) using UV lamp emitting at 254 nm and VUV lamp emitting at both 254 and 185 nm. VUV irradiation was reported to be ef?cient for the degradation of organics. This effect was attributed to the additional photons of 185 nm and the higher energy of illumination

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with VUV lamps than with UV lamps. The order of TOC removal rate of each phenol was shown to be VUVPCD N VUVPD N UVPCD N UVPD, except for 4-NP under VUV irradiation which had no signi?cant difference due to TiO2 catalyst. The degradation rates in both processes are reported to follow the order: 4-CP N hydroquinoneN 4-NP. 3.3. Initial concentration Successful application of the photocatalytic oxidation system requires the investigation of the dependence of photocatalytic degradation rate on the substrate concentration. Hong et al. [56] examined the effect of initial concentration (50–400 ppm) on the degradation of phenol over TiO2 when the ratio of H2O/titanium isopropoxide is 75. Highest degradation was observed at 50ppm initial concentration. Parida and Parija [57] studied the effect of initial substrate concentration (2–25 g/l) on the photocatalytic degradation of phenol under sunlight, visible and UV light, respectively. With the increase in the substrate concentration, the degradation ef?ciency decreased from 100% to 60% under solar irradiation. Under UV light, the degradation was found to decrease from 94% to 52% with increasing initial concentration. The degradation was reported to decrease from 95% to 50% under visible light. Shukla et al. [34] studied the effect of initial concentration (12.5–37.5 ppm) on the degradation of phenol using UV–vis/ZnO/persulfate. Under the conditions tested (ZnO = 0.4 g/l, persulfate = 2 g/l and power = 330W), the degradation was highest at 12.5 ppm. Parida et al. [58] examined the effect of substrate concentration on the degradation of 4-nitrophenol (20–100 mg/l). The amount of degradation was found to decrease from 100 to 40.9% as the initial substrate concentration increases from 20 to 100 mg/l. This is due to the absorption of light by the substrate at high concentration for the given catalyst loading. Priya and Madras [33] examined the effect of initial concentration (10–76 ppm) on the photocatalytic degradation of 2,4-dinitrophenol in the presence of combustion synthesised TiO2 and P25. Maximum degradation was observed at 76 ppm. Pardeshi and Patil [59] reported that the degradation ef?ciency of phenol was shown to decrease as the phenol concentration increases from 25 to 300 mg/l in the presence of ZnO and sunlight. When the concentrations of phenol are high, more and more phenol molecules are adsorbed on the surface of ZnO. In contrast, the relative number of ·OH and ·O2? radicals attacking the phenol molecules decreases due to constant reaction conditions. As a result the photocatalytic degradation decreases. Venkatachalam et al. [53] examined the effect of initial concentration (50–300 mg/l) on the photocatalytic degradation of 4-CP using nano TiO2. The degradation rate was found to increase up to an initial concentration of 250 mg/l and then decreased. The degradation follows pseudo-?rst order kinetics at low 4-CP. This is ascribed to the screening effect at concentrations greater than 250 mg/l, and hence degradation ef?ciency decreased. Lathasree et al. [60] studied the effect of initial concentration (40–100 ppm) on the photocatalytic degradation of phenol using ZnO as catalyst. The initial rates of photodegradation were high at the lower concentration range and it decreased with increase in concentration. The degradation was found to follow the ?rst order kinetics. For chlorophenols, the initial rate was observed to increase with increase in the initial concentration range 40–60 ppm and decreased with further increase in concentration. Barakat et al. [61] tested the effect of initial concentration (12.5–75 ppm) on the photocatalytic degradation of 2-CP using 0.036 mol% Co-doped TiO2. At a concentration of 12.5 ppm, 2-CP can be totally decomposed within 3 h. The degradation process follows a pseudo ?rst-order reaction. The observed rate constant was shown to vary from 0.020 to 0.008 min? 1 as the concentration increases from 12.5 to 75 mg/l. Kusvuran et al. [37] examined the effect of initial concentration (0.1–0.5 mM) for the photocatalytic degradation of di and tri-substituted phenols using UV/TiO2. The degradation rate constant of TCP was increased from 0.0372 to 0.0136 min? 1 under the conditions tested. For

TBP and TMP, the degradation rate constant was shown to increase from 0.0252 to 0.0071 min? 1 and 0.0153 to 0.0034 min? 1 respectively. TCP with the strongest electronegativity character of the compounds studied exhibits lower af?nity to be adsorbed on TiO2 catalyst. As a result of that, the number of reactive sites on TiO2 increased and higher degradation levels were observed when initial TCP concentration was low. Chiou et al. [62] tested the effect of initial concentration (12.5–200 mg/l) on the photocatalytic degradation of phenol using 0.072 mol% Pr-doped TiO2. The highest degradation was 100% after 60 min irradiation for a range of concentration 12.5–25 mg/l. The rate of degradation was shown to decrease as the phenol concentration increases from 50 to 200 mg/l. Chiou et al. [63] studied the effect of initial concentration (15.3–83.5 mg/l) on the photocatalytic degradation of phenol and m-nitrophenol by UV/TiO2 process. More than 90% of phenol is degraded within 60 min irradiation when the initial concentration is 15.3 mg/l. However, 42% phenol is removed in 180 min when initial concentration increases up to 83.5 mg/l. Under the identical conditions, 83% and 43% of m-NP is degraded at an initial concentration of 15.3 and 83.5 mg/l, respectively. The limited number of surface sites on TiO2 particles may control the photodegradation. As indicated in several investigations, as the concentration of target pollutant increase, more and more molecules of the compound get adsorbed on the surface of the photocatalyst. Therefore, the requirement of reactive species (·OH and ·O?) needed for the degradation 2 of pollutant also increases. However, the formation of ·OH and ·O? 2 on the catalyst surface remains constant for a given light intensity, catalyst amount and duration of irradiation. Hence, the available OH radicals are inadequate for the pollutant degradation at higher concentrations. Consequently the degradation rate of pollutant decreases as the concentration increases [25]. In addition, an increase in substrate concentration can lead to the generation of intermediates which may adsorb on the surface of the catalyst. Slow diffusion of the generated intermediates from the catalyst surface can result in the deactivation of active sites of the photocatalyst, and consequently result in a reduction in the degradation rate. In contrast, at low concentration the number of catalytic sites will not be a limiting factor, and the rate of degradation is proportional to the substrate concentration in accordance with apparent ?rst-order kinetics [48,49]. Several investigations adequately described the dependence of photocatalytic degradation rates on the concentration of various phenols and dyes by the Langmuir–Hinshelwood (L–H) kinetics model [64–66]. The L–H model is used to describe the dependence of the observed reaction rate on the initial solute concentrations [67], although the model parameters will be strongly dependent on the composition of the ef?uent and other reactor operating conditions. 3.4. Catalyst loading Several studies have indicated that the photocatalytic rate initially increases with catalyst loading and then decreases at high values because of light scattering and screening effects. The tendency toward agglomeration (particle–particle interaction) also increases at high solids concentration, resulting in a reduction in surface area available for light absorption and hence a drop in photocatalytic degradation rate. Although the number of active sites in solution will increase with catalyst loading, a point appears to be reached where light penetration is compromised because of excessive particle concentration. The trade off between these two opposing phenomena results in an optimum catalyst loading for the photocatalytic reaction [15]. A further increase in catalyst loading beyond the optimum will result in non-uniform light intensity distribution, so that the reaction rate would indeed be lower with increased catalyst dosage. Table 4 presents the in?uence of catalyst loading on the photocatalytic degradation of phenols. Parida et al. [58] studied the effect of ZnO concentration (0.2 to 2.0 g/l) on the photocatalytic degradation of 4-nitrophenol under solar irradiation. At high concentration, the

S. Ahmed et al. / Desalination 261 (2010) 3–18 Table 4 Effect of catalyst loading on the photocatalytic degradation of various phenols. Pollutant 4-Nitrophenol Phenol 4-Chlorophenol Phenol 2-Chlorophenol Phenol Phenol Light source Solar Solar UV UV UV UV UV Photocatalyst ZnO ZnO TiO2 ZnO Co-TiO2 TiO2 Pr–TiO2 Tested catalyst loading (g/l) 0.2–2.0 0.5–3.5 1.0–5.0 1.0–3.0 0..005–0.03 1.0–4.0 0.–1.2 Optimum loading (g/l) 0.6 2.5 2.0 2.0 0.01 2.0 1.0 Ref. [58] [59] [53] [60] [61] [56] [62]

9

degradation rate was observed to have levelled off. Under the conditions tested, the degradation of 4-NP was found to increase with increase in ZnO loading up to 0.6 g/l. An increase in the amount of catalyst provides an increased number of active sites for adsorption; however, the simultaneous increase in solution opacity causes a decrease in the penetration of the photon ?ux. This suggests that the amount of photocatalyst to be used should maintain a balance between these two opposing effects. Selvam et al. [40] indicated that the degradation rate constant of 4-?uorophenol increases from 0.0152 to 0.0358 min? 1 as the concentration of TiO2 increases from 50 to 150 mg in the presence TiO2 P25. Further increase in catalyst loading (150 to 250 mg) results in a decrease in the rate constant from 0.0358 to 0.0296 min? 1. In the case of ZnO, a similar observation has been made. This was related to the aggregation of free catalyst particles and the screening effect resulting from the excessive opacity of the solution. The optimum concentrations were found to be 150 mg TiO2 and 200 mg ZnO for ef?cient removal of 4-?uorophenol. The photodegradation ef?ciency of BPA was shown to increase when the dosage increased up to 100 mg/200 cm3, and then the ef?ciency was nearly 100% in the TiO2 dosage range 100–600/200 cm3 [68]. Pardeshi and Patil [59] studied the effect of ZnO concentration (0.5 to 3.5 g/l) on the solar photocatalytic degradation of phenol. As the amount of ZnO increases, the number of phenol molecules adsorbed was also increased due to an increase in the number of ZnO particles. Thus the photocatalytic degradation was enhanced. Further increase in the amount of catalyst showed an adverse effect. Under the conditions tested, the highest degradation was observed at 2.5 g/l of ZnO. The decrease in degradation beyond 2.5 g/l is attributed to the screening effect of excess ZnO particles in the solution and scattering of light. Shukla et al. [34] observed an optimum concentration of 0.8 g/l of ZnO for the solar photocatalytic degradation of phenol under a UV–vis/ZnO/persulfate system. For ZnO loading of 0.2 g/L, the degradation ef?ciency of phenol is shown to be 95% at 4 h while a complete degradation is reached within 3 h for 0.4 g/L or within 2 h at higher amounts of catalyst loading. In contrast to accelerating the rate of reaction resulting from the addition of excess catalyst, it can possibly cause a negative effect by reducing the light transmittivity due to the formation of an opaque solution. Venkatachalam et al. [53] tested the effect of catalyst loading (1–5 g/l) on the photocatalytic degradation of 4-CP and observed an optimum catalyst concentration

of 2.0 g/l for the degradation of 4-CP. At high catalyst loading, the reported low degradation rate is attributed to the deactivation of activated molecules through collision with ground state titania molecules. Lathasree et al. [60] examined the effect of ZnO on the photocatalytic degradation of phenols in the range of 1.0–3.0 g/l. The optimum degradation was shown to be at 2.0 g/l ZnO under the experimental conditions. The decrease in the degradation rate beyond catalyst loading of 2.0 g/l was attributed to the screening effect of excess catalyst particles in the solution. Barakat et al. [61] studied the effect of 0.036 mol% Co-doped TiO2 concentrations (5–30 mg/L) on the photocatalytic degradation of 2-CP. The degradation reached a maximum value of 93.4% with catalyst dosage of 10 mg/L. However, a further increase in the catalyst dosage slightly decreased the degradation ef?ciency. Hong et al. [56] investigated the effect of catalyst dosage (1.0–4.0 g/l) on the photocatalytic degradation of phenol. Under the conditions tested, optimum degradation was achieved at 2.0 g/l. Chiou et al. [62] studied the effect of 0.072 mol% Pr-doped TiO2 concentration (0.2–1.2 g/L) on the photocatalytic degradation of 50 mg/l phenol. The degradation increases with increasing dosage of 0.072 mol% Pr-doped TiO2, and reaches a maximum of 96.5% at a dosage of 1.0 g/L. However, a further increase in catalyst dosage to 1.2 g/L, slightly reduces the photodegradation ef?ciency due to the decrease in light penetration by the suspension. In order to ensure uniform light intensity in the photocatalytic reactor, optimum catalyst loading must be determined. 3.5. Medium pH Organic compounds in wastewater differ greatly in several parameters, particularly in their speciation behaviour, solubility in water and hydrophobicity. While some compounds are uncharged at common pH conditions typical of natural water or wastewater, other compounds exhibit a wide variation in speciation (or charge) and physico-chemical properties. At a pH below its pKa value, an organic compound exists as a neutral species. Above this pKa value, an organic compound attains a negative charge. Some compounds can exist in positive, neutral, as well as negative forms in aqueous solution. This variation can also signi?cantly in?uence their photocatalytic degradation behaviour. The pH of the wastewater can vary signi?cantly. The pH of an aquatic environment plays an important role on the

Table 5 In?uence of pH on the photocatalytic degradation of phenol and substitute phenols. Pollutant 4-Nitrophenol Phenol 4-Chlorophenol 2-phenyl phenol 2-Chlorophenol 4-?urophenol Phenol m-Nitrophenol 4-Chlorophenol
a

Light source Solar Solar UV UV UV UV UV UV UV

Photocatalyst ZnO ZnO TiO2 ZnO Co–TiO2 ZnO/TiO2 TiO2 TiO2 N–TiO2

Tested pH range 2.0–7.0 3.0–11.0 4.0–9.0 6.6–12.0 4.0–12.0 4.0–9.0 3.0–8.0 4.1–12.7 2.0–5.0

Optimum pH 6.0 5.0–7.0 5.0 12.0 9.0 7.0 5.0 8.9 3.0a

Ref. [58] [59] [53] [73] [61] [40] [74] [63] [75]

Precursor solution pH.

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photocatalytic degradation of organic contaminants since it determines the surface charge of the photocatalyst and the size of aggregates it forms [69,70]. The surface charge of a photocatalyst and ionization or speciation (pKa) of an organic pollutant can be profoundly affected by the solution pH. Electrostatic interaction between semiconductor surface, solvent molecules, substrate and charged radicals formed during photocatalytic oxidation is strongly dependent on the pH of the solution. In addition, protonation and deprotonation of the organic pollutants can take place depending on the solution pH. Sometimes protonated products are more stable under UV-radiation than their main structures [71]. Therefore the pH of the solution can play a key role in the adsorption and photocatalytic oxidation of pollutants. The ionization state of the surface of the photocatalyst can also be protonated and deprotonated under acidic and alkaline conditions respectively as shown in the following reactions: pHbPzc TiOH ? H →TiOH2
? ? ?

?5:1? ?5:2?

pHNPzc TiOH ? OH →TiO ? H2 O

?

The point of zero charge (Pzc) of the TiO2 (Degussa P25) is widely investigated/reported at pH ? 6.25 [25]. While under acidic conditions, the positive charge of the TiO2 surface increases as the pH decreases (Eq. (5.1)); above pH 6.25 the negative charge at the surface of the TiO2 increases with increasing pH. Table 5 shows the in?uence of pH on the photocatalytic degradation of various phenols. At pH 5–9, the Zn2+ and Fe3+ co-doped TiO2 is found to show higher photoactivity for the degradation of phenol, whereas an optimal pH range of 3–4 was reported for the pure TiO2 [72]. Venkatachalam et al. [53] studied the effect of pH (4–9) on the photocatalytic degradation of 4-CP (250 mg/l) using nano TiO2. The rate of degradation in the acidic pH range (at 5) was found to be higher than the alkaline pH. The high degradation rate in the acidic pH is due to enhanced adsorption of 4CP on the surface of nano TiO2 that carries positive charge. In the acidic pH, minimisation of electron–hole recombination is also an important factor for the enhanced degradation of 4-CP. In alkaline pH (at 9), the degradation rate was found to be less due to the negative charge of nano TiO2 and 4-CP. Parida et al. [58] examined the effect of pH (2–7) on the photocatalytic degradation of 4-NP. The degradation rate was found to increase 92% at pH 6.0 and thereafter decreased with increasing pH of the solution. At pH 6, the surface charge of ZnO can attract the largest amount of negatively charged anions of 4-nitrophenol. At pH above 7, negative charges predominate on the surface of ZnO which hinders the degradation of 4-NP. Pardeshi and Patil [59] investigated the effect of pH (3–11) on the solar photocatalytic degradation of phenol using ZnO suspension. Under the conditions tested, the highest degradation was achieved at pH 5–7. At pH 5–7, most of the phenol remains undissociated, hence the maximum number of phenol molecules are adsorbed on the ZnO surface and consequently result in enhanced photodegradation. In an alkaline medium, the ZnO surface is negatively charged and phenolate intermediates may be repelled away from the ZnO surface which opposes the adsorption of substrate molecules on the surface of the catalyst. As a result, phenol degradation decreases in alkaline medium. Using ZnO, Lathasree et al. [60] studied the effect of pH (3.5–9) on the photocatalytic degradation of phenol, o-chlorophenol and p-chlorophenol. The phenol degradation was reported to be favourable in mild acidic and neutral solutions. However, o-chlorophenol and p-chlorophenol was found to undergo degradation at a faster rate at lower pH values. This behaviour is attributed to the undissociated state of the chlorophenol species in acidic conditions which favours its adsorption on the ZnO surface. At high pH, chlorophenols exist as negatively charged anions which weaken its adsorption due to electrostatic repulsion between catalyst and anions.

Khodja et al. [73] examined the effect of pH (6.6–12) on the photocatalytic degradation of 2-phenyl phenol using ZnO. The degradation ef?ciency was shown to be greater at higher pH due to the abundance of hydroxyl radicals or to the anionic form of the tested compound to be oxidized compared to its molecular form. Barakat et al. [61] investigated the effect of pH (4–12) on the photocatalytic degradation of 2-CP using 0.036 mol% Co-doped TiO2. At lower pH (4–7), the degradation ef?ciency was signi?cantly less as compared to higher pH, with a maximum value of 93.4% at pH 9. They reported that higher concentrations of hydroxyl ions are available at higher pH values to react with the holes to form hydroxyl radicals (·OH), thereby enhancing the photodegradation of 2-CP. Further increase in pH indicated only a small increase in the photodegradation ef?ciency. Selvam et al. [40] studied the effect of pH (4–7) on the photocatalytic degradation of 4-FP. After 30 min irradiation, the pseudo-?rst order rate constants were 0.0233, 0.0370 and 0.028 min? 1 for P25 and 0.0213, 0.0325 and 0.0231 min? 1 for ZnO at pH 4, 7 and 9 respectively. Since the pKa of 4-?urophenol is 9.89, it exists as molecular form for the tested pH range. As a result, higher adsorption and consequently higher degradation are reported to be observed at neutral pH for both catalysts. In the presence of P25 and ZnO, the effect of pH on the photocatalytic degradation of 4-?uorophenol was observed to be pH 7 N pH 9 N pH 4. Akbal and Onar [74] examined the effect of pH (3–8) on the photocatalytic degradation of phenol. At low pH degradation of phenol was rapid, and the highest degradation (70.3%) was achieved at pH 5. The addition of H2O2 to the UV/TiO2 system was observed to enhance the phenol degradation to 99.2%. Chiou et al. [63] studied the effect of pH (4.1–12.7) on the photocatalytic degradation of phenol and m-NP (C0 = 0.51 mM, TiO2 = 1 g/l). The degradation ef?ciency is above 65% within 180 min with the pH lower than 11; however, it reaches 53% at pH 12.7. The degradation is over 85% within 180 min for m-NP, and the degradation was more effective in neutral and weakly alkaline media. The degradation decreases in the order pH 8.9 N pH 7 N pH 11. In neutral to weak alkaline media, TiO– is the predominant group on TiO2 particles, and phenols are primarily present in the molecular forms. Both of them combined by hydrogen bonding simply, and thus increased the amount of adsorption and enhanced the photodegradation rate. Yu et al. [75] tested the in?uence of a precursor solution pH (2.0–5.0) on the photocatalytic activity of N–TiO2 calcined at 500 °C for the degradation of 4-CP. The ef?ciency of N-doped TiO2 was found to increase as the pH decreased from 5.0 to 3.0. Further decrease in pH to 2.0 affected the photocatalytic activity of the catalyst negatively. Hence, the optimum pH for that particular catalyst's preparation was 3.0. A possible reason for the adverse effect on the photocatalyst performance at low pH is that the increased of H+ concentration may restrain hydrolysation of Ti(OBu)4 and thereby reduce the crystal size of the prepared N-doped TiO2 nanoparticles. In sum, different phenols have different activity in photocatalytic reaction. Some are degraded faster in acidic pH, while others degrade faster at higher pH. All these effects can be attributed to the nature of the pollutant to be degraded. Thus, it is important to study the nature of the pollutants to be degraded, and determine the optimum pH. 3.6. Co-occurring substances The amount of UV absorption is in?uenced by water transmittance over the spectral UV range of interest. Some common constituents that affect water transmittance are dissolved organic matter, nitrate and ferrous/ferric ions. The presence of “spectator's components” in water can affect adversely the contaminant degradation rates. Inorganic anions, such as phosphate, sulphate, nitrate, and chloride, have been reported to limit the performance of solar based photocatalysis [76]. Bicarbonate in particular is detrimental to reactor performance as it acts as a hydroxyl radical scavenger [76]. Long time experience with photocatalytic oxidation systems showed that humic substances in contaminated water can strongly adsorb titanium

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dioxide particles and reduce activity toward the target substances [77]. Lin and Lin [78] observed that the presence of humic acid caused a signi?cant retardation on the photocatalytic degradation of 4-chlorophenol. The observed retardations of humic acids were related to the inhibition (surface deactivation), competition and light attenuation effects. Moreover, the presence of humic acid in the reaction mixture has been reported to signi?cantly reduce light transmission, and therefore the photooxidation rate. Humic acid can also compete with organohalides for the active sites on the TiO2 surface. Popadam et al. [79] examined the effect of the water matrix on the decolourisation of acid orange 20. The presence of 5 g/l NaCl and Na2SO4 led to decreased decolourisation which was attributed to the trapping of photo generated valence band holes and the hydroxyl radicals by the respective anions. Aarthi and Madras [80] investigated the effect of metal ions (Cu2+,Fe3+,Zn2+, and Al3+) on the photocatalytic degradation of Rhodamine B which was shown to follow the order of Cu2+ N Al3 N Zn2 N Fe3+. In the presence of Cu2+ and Zn2+, the degradation of Malachite Green was reduced by 76 and 66% respectively. In the presence of metal ions, the surface of TiO2 is modi?ed due to the adsorption of cations or anions. The effect of metal ions on the photocatalytic degradation rates is attributed to the suppression of OH radicals due to trapping of the conduction band electrons by the adsorbed metal ions. Khodja et al. [73] examined the effect of sodium chloride, sulphate and nitrate concentration on the photocatalytic degradation of 2-phenyl phenol. A large amount of these anions is shown to affect negatively the degradation rate. This inhibition can be attributed to a competitive adsorption on the ZnO of these anions and molecules of substrate. Selvam et al. [40] examined the effect of transition metal ions on the photocatalytic degradation of 4-?urophenol which was shown to be in the order of Mg2+ N Fe2+ N Fe3+ N Cu2+ in an acidic medium, whereas the trend was observed to be Mg2+ N Fe3+ N Cu2+ N Fe2+ in a basic medium. The enhancement of degradation in the presence of these ions except Cu2+ is due to the increase of charge separation by accepting the conduction band electrons. The inhibition of inorganic anions on the degradation of 4-FP was demonstrated to be CO2? N HCO? N Cl? N NO? N SO2?. The main 3 3 3 4 inhibition from these anions is attributed to their adsorption on the TiO2 surface. Addition of CO2?, HCO? and Cl? ions decreases degradation of 3 3 4-?urophenol due to the hydroxyl radical quenching by these ions. The · SO2? ions react with hydroxyl radicals to form SO4? ions. The sulphate 4 radical anion can accelerate the reaction, and no quenching was observed for the NO? ion. Naeem and Feng [81] studied the inhibition 3 effect of anions on the photocatalytic degradation of phenol. The order of effects was demonstrated to be Cl? N SO2? N NO? N CO2?. This is due to 4 3 3 the fact that the Cl? ions act as radical scavengers, and can react with hydroxyl radicals to form ·Cl2 or ·ClOH? which may retard the photocatalytic oxidation reaction dramatically. But the inhibitive effect of SO2? can be attributed to the competitive adsorption of SO2? and the 4 4 target organic compound on the catalyst surface. At certain concentrations, Fe3+ is reported to improve the degradation rate whereas Ca2+, Cu2+, and Mg2+ hinder the degradation process. The improvement in degradation with Fe3+ addition is due to the electron scavenging effect that prevents the recombination of electrons and holes which favours · the formation of ·OH and O2? on the TiO2 surface. The decrease in degradation in the presence of other metal ions is attributed to low reduction potential. The effect of surfactants (0.05 mM/L) on the degradation rate of phenol was reported to be sodium dodecyl benzene sulphate (SDBS) N sodium dodecyl sulfonate (AS) N sodium dodecyl sulphate (SDS). The presence of anions is reported to alter the ionic strength of the solution, therefore in?uencing the catalytic activity and hence the photocatalytic degradation. Organic solvents are mostly present in many industrial wastewaters. As a result, their effects on the photocatalytic degradation system have been investigated in several studies. Lin and Lin [78] examined the effects of acetonitrile or isopropanol on the photo-

catalytic oxidation of 4-Chlorophenol. The degradation rate of 4Chlorophenol was shown to decrease from 1.0436 h? 1 to 0.0525 h? 1 as the acetonitrile content increases in the reaction mixture. A similar retardation effect was also observed when 5% isopropanol was added into the acetonitrile: water (1:1) reaction mixture. Daneshvar et al. [82] studied the effect of ethanol (0–9% v/v) addition on the photodegradation of Acid Orange 7 in the presence of ZnO. Under the conditions examined, the addition of ethanol inhibited the degradation of AO7 due to ·OH competitive reactions with AO7 and ethanol. Thus, the incorporation of a pre-treatment process has been indicated to have a bene?cial effect on the performance of photocatalytic oxidation. 3.7. Oxidants/electron acceptor The electron/hole recombination is one of the main drawbacks in the application of TiO2 photocatalysis as it causes a waste of energy. In the absence of suitable electron acceptor or donor, recombination step is predominant and thus it limits the quantum yield. Thus it is crucial to prevent electron–hole recombination to ensure ef?cient photocatalysis. Molecular oxygen is generally used as an electron acceptor in heterogeneous photocatalytic reactions. Addition of external oxidant/electron acceptors into a semiconductor suspension has been shown to improve the photocatalytic degradation of organic contaminants by (1) removing the electron–hole recombination by accepting the conduction band electron; (2) increasing the hydroxyl radical concentration and oxidation rate of intermediate compounds; and (3) generating more radicals and other oxidizing species to accelerate the degradation ef?ciency of intermediate compounds [25–32]. Since hydroxyl radicals appear to play an important role in the photocatalytic degradation, several researchers have investigated the effect of the addition of electron acceptors such as H2O2, KBrO3, and (NH4)2S2O8 on the photocatalytic degradation of various organic compounds [26–32] to enhance the formation of hydroxyl radicals as well as to inhibit the electron/hole (e?/h+) pair recombination. In all cases, the addition of oxidants has resulted in higher pollutant degradation rates compared to molecular oxygen. In most of the cases, the order of enhancement is reported to be UV/TiO2/H2O2 N UV/TiO2/ BrO? N UV/TiO2/S2O2?. The enhancement of degradation rate is due 3 8 to the reaction between BrO3– and conduction band electrons. This reaction reduces the recombination of electron–hole pairs. BrO3 ? 6eCB ? 6H →Br ? 3H2 O
? ? ? ?

?7:1?

S2O2? can generate sulphate radical anion (SO.4?) both thermally 8 and photolytically in aqueous solution. SO.? then reacts with H2O to 4 produce .OH radicals. S2 O8 →2SO4
:? 2? :?

?7:2?
2? ?

SO4 ? H2 O→ OH ? SO4 ? H

:

?7:3?

With the addition of H2O2, the enhancement of degradation is due to the increase in the hydroxyl radical concentration as shown by Eqs. (7.4) and (7.5); H2 O2 ? eCB → OH ? OH H2 O2 → OH ? OH
: : ? : ?

?7:4? ?7:5?

The degradation ef?ciency of the UV/TiO2/oxidant process is slightly more in an acidic medium than in a basic medium. Selvam et al. [40] studied the effect of oxidants on the photocatalytic degradation of 4-?urophenol which was shown to be in the order of IO? N BrO? N S2O8? N H2O2 N ClO?. After 15 min irradiation at pH 4, the 4 3 2 3

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degradations of 4-?ourophenol were observed to be 79.86%, 66.89%, 60.88%, 53.85% and 46.35% in UV/TiO2/IO?, UV/TiO2/BrO?, UV/TiO2/ 4 3 H2O2, UV/TiO2/S2O2? and UV/TiO2/ClO? processes respectively. IO? 8 3 4 was found to be the most ef?cient oxidant due to the generation of a number of highly reactive intermediate radicals such as IO.3, OH. and IO?. The same trend is also observed in both acidic and basic mediums 4 for the UV/ZnO/oxidant process. However, the UV/TiO2/oxidant process is slightly more ef?cient than the UV/ZnO/oxidant process. The higher ef?ciency of BrO? than S2O2? is due to the electron 3 8 scavenging effect of BrO?. Under visible light irradiation, the in?uence 3 of oxidants on the degradation of Acid Red 88 by Ag–TiO2 was found to be peroxomonosulfate N peroxodisulfate N hydrogen peroxide [83]. The addition of 50 ml H2O2 to a 1.00 mM p-dihydroxybenzene solution was reported to increase the removal ef?ciency from 37 to 80% after 200 min reaction time due to the high oxidising capacity of hydroxyl radicals [84]. Chiou et al. [54] investigated the effect of H2O2 addition from 1.77 to 88.2 mM on the photocatalytic degradation of phenol. The addition of H2O2 from 1.77 to 8.82 mM leads to an increase in removal ef?ciency from 58 to 84% within 3 h. In contrast, phenol is completely degraded within 2.5 and 1 h when the level of H2O2 increases to 44.1 and 88.2 mM respectively. 3.8. Calcination temperature Structure and size of TiO2 crystallites are signi?cantly dependent on calcination temperatures. Thermal treatment of TiO2 gels at higher temperatures promotes phase transformation from thermodynamically metastable anatase to the more stable and condensed rutile phase. As the dehydration process occurs during heat treatment, crystallites grow to dimensions larger than those of the original particles. Table 6 shows the in?uence of calcination temperatures on the photocatalytic degradation of various phenolic compounds. Nanocrystalline titanium oxide powder for catalytic purposes can be applied by a variety of techniques including sol–gel method, solvothermal process, reverse micellar, hydrothermal method and electrochemical methods. An et al. [85] studied the effect of calcination temperatures (350–650 °C) on the photocatalytic degradation of 2,4,6-tribromophenol (TBP) using mesoporous-TiO2 prepared by hydrothermal assistant sol–gel method. The TBP degradation rate constant was found to increase from 0.010 to 0.018 min? 1 when calcination temperature was increased from 350 to 550 °C. A sharp decrease in the rate constant to 0.004 min? 1 was observed when the sample was thermally treated at 650 °C. As the treatment temperature increases from 350 to 550 °C, the resulting TiO2 increases its crystallite size from 9.8 to 18.4 nm, while BET surface area deceased from 142 to 49 m2/g. The increased photocatalytic degradation ef?ciency was attributed to the increased crystallinity. The decrease in photocatalytic degradation ef?ciency at 650 °C was ascribed to the completely collapsed mesoporous structure of TiO2. Pecchi et al. [86] investigated the effect of calcination temperatures on the photocatalytic degradation of pentachlorophenol using the sol–gel method under the pH

range 3–9. The degradation rate constant was shown to increase from 6.9 × 103 to 7.9 × 10? 3 min? 1 and 7.0 × 10? 3 to 10.1 × 10? 3 min? 1 respectively for pH values 3 and 5 as the calcination temperatures increased from 300 to 500 °C. Nonetheless, the rate constant was higher at 300 °C and 500 °C compared that for 400 °C at pH 9. Increasing calcination temperature results in reduced surface area from 175 to 67 m2/g at pH 3 and 170 to 68 m2/g at pH 5, and increased crystallinity and particle size from 3.3 to 8.7 nm at pH 3 and 4.0 to 9.8 nm at pH 9. No signi?cant changes in the anatase and rutile proportion are observed as the calcination temperature increases. Gorska et al. [87] studied the effect of calcination temperatures (350 °C–750 °C) on the photocatalytic degradation of 0.21 mM phenol under UV (25 b λ b 400 nm) and visible (λ N 400 nm) light irradiation. After 60 min irradiation, the maximum amount of phenol degradation was about 80% when P25 TiO2 was calcinated at 350 °C in the visible light. In addition to the largest BET surface area (205.8 m2/g) and the smallest crystallite size (8.4 nm) and average pore diameter (8.3 nm) of TiO2, this effect was attributed to the ef?cient absorption of light vis region due to the existence of the largest amount of carbon in aromatic C–C bonds (10.1 at.%). However, the photocatalyst calcinated at 450 °C was found to have the highest activity in UV light irradiation. Silveyra et al. [88] tested the effect of calcination temperatures (600–800 °C) on the photocatalytic degradation of phenol using N-doped TiO2 P-25. Highest degradation was obtained at 600 °C and no variation was observed at 700 °C and 800 °C. No discernible change in anatase to rutile ratio is noticed for N2-doped TiO2 at 600 °C whereas this ratio is beginning to be noticeable at 700 to 800 °C. Liqiang et al. [89] observed that the rutile phase began to appear when the pure TiO2 samples calcinated at 600 °C, and the anatase phase disappeared when the calcination temperature was at 700 °C. In contrast a little rutile phase appeared in the La-doped TiO2 samples when the sample was calcinated at 700 °C, and even most was shown to be anatase phase at a calcination temperature of 800 °C. A certain temperature, 600 °C was reported to be the optimum calcination temperature for the highest photocatalytic degradation of phenol using 5 mol% La-doped TiO2. Wawrzyniak et al. [90] studied the effect of calcination temperatures (100–800 °C) on the photocatalytic degradation of phenol and two azo dyes (reactive Red, direct Green) by N doped-TiO2 under visible light irradiation. The highest effectiveness of phenol degradation was achieved at a calcination temperature of 700 °C. In the case of degradation of both of the azo dyes, the highest degradation was observed at 500 °C and 600 °C. Cheng et al. [95] studied the effect of calcination times (0.5–3 h) on the photocatalytic degradation of 4-CP using carbon modi?ed TiO2 calcinated at 350 °C under visible light irradiation. The surface area of synthesised TiO2 decreases from 257 to 119 m2/g and the crystallite size increases from 4.35 to 5.58 nm as the calcination times increase from 0.5 to 3.0 h. The catalyst calcinated for 1 h is shown to have higher photocatalytic activity than the catalyst calcinated for 0.5 h. The higher activity of the catalyst calcinated for 1 h is due to its enhanced crystallization in anatase TiO 2 which acts as the

Table 6 In?uence of calcinations temperature on the photocatalytic degradation of phenols. Pollutant Phenol Phenol Phenol Phenol p-chlorophenol 4-CP Phenol Phenol 2,4,6-TBP Phenol Light source UV Visible Visible UV Solar UV UV UV/Visible Visible UV UV Photocatalyst TiO2 P-25 TiO2 P-25 N2–TiO2 La– TiO2 Au–TiO2 TiO2 Zr4+–TiO2 I–TiO2 N2–TiO2 TiO2 Pr–TiO2 Range of calcination temperature 350–750 °C 350–750 °C 600–800 °C 600–800 °C 275–500 °C 300–500 °C 105–1100 °C 400–600 °C 350–500 °C 350–650 °C 100–800 °C Optimum calcination temperature 350 °C 450 °C 700 °C 600 °C 275 °C 500 °C 900 °C 400 °C 400 °C 550 °C 600 °C Ref. [87] [87] [88] [89] [91] [86] [92] [93] [94] [85] [62]

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recombination centres for photogenerated electrons and holes. Nonetheless, calcination time of more than 1 h results in reduced photocatalytic activity due to the decrease of carbonaceous residues and surface area. Shaktivel et al. [96] studied the effect of calcination temperatures (400–600 °C) on the photocatalytic degradation of 4-CP under visible light irradiation (λ ≥ 455 nm) using a nitrogen-doped TiO2 prepared by TiCl4 hydrolysis in tetrabutylammonium hydroxide. A sample calcinated at 400 °C for 1 h is reported to show 70% TOC reduction of 0.25 mM 4-CP. Porter et al. [97] examined the microstructural changes in Degussa P-25 due to heat treatment. The TiO2 powder was annealed from 600 to 1000 °C. Under UV irradiation, the apparent crystallite size and rutile content of the catalyst increased with increasing calcination temperatures, whereas the speci?c surface area and the rate of phenol degradation was reported to decrease. Sonawane and Dongare [91] tested the effect of calcination temperatures (275–500 °C) on the photocatalytic activity of 2%-Au/TiO2 for the degradation of phenol under solar light. The catalyst calcinated at 275 °C exhibited the highest activity compared to the samples calcinated at higher temperatures. This effect is attributed to the formation of the most active pure anatase phase at this temperature and the particle size. At lower temperatures the particle growth will be limited to the surface area and the number of active sites will be more, hence the photocatalytic activity obtained is more than for prepared Au/TiO2 and catalyst calcinated at higher temperatures. Zhang et al. [43] studied the effect of calcination temperatures (300–900 °C) on the photocatalytic degradation of MO using ZnO– SnO2. At 350 °C, the higher photocatalytic activity of coupled oxides was reported due to variation in phase composition and particle size. Using Z2S as the catalyst, an optimum temperature of 600 °C was found for the degradation of MO [44]. Lukac et al. [92] studied the effect of annealing temperature (105–1100 °C) on the photocatalytic activity of Zr-doped TiO2 for the degradation of 4-chlorophenol. The catalysts annealed between 800 and 900 °C are reported to be more ef?cient compared to the standard Degussa P25. The optimum size of the Zr-doped anatase crystallites is 69.2 and 86.5 nm obtained at 875 and 900 °C respectively. The photocatalyst annealed at 900 °C was reported to have a 1.5 times higher degradation rate than the standard TiO2 P25. The most photoactive catalyst annealed at 900 °C was shown to have 87 wt.% of anatase and 13 wt.% of rutile. Barakat et al. [61] studied the effect of annealing temperatures (100–800 °C) on the ef?ciency of 0.036 mol% co-doped TiO2 for the photocatalytic degradation of 2-CP. The sample annealed at 800 °C is found to have rutile structure and shows lower degradation ef?ciency compared to the one annealed at 600 °C with anatase phase. The enhanced photoreactivity is attributed to the small particle size of TiO2 (about 25 nm) with a large surface area. The band gap energy measured for this anatase sample was 3.2 eV, which is identical to that of the bulk TiO2. Hong et al. [93] examined the effect of calcination temperatures (400–600 °C) on the photocatalytic degradation of phenol under UV

and visible light irradiation (λ N 400 nm) using I-doped TiO2 with anatase phase. Of the temperatures tested, I-doped TiO2 calcinated at 400 °C shows signi?cantly higher photocatalytic activity compared to undoped TiO2 nanoparticles (P-25 and pure TiO2). Wang et al. [94] tested the effect of calcination temperatures (350–500 °C) on the photocatalytic activity of N2-doped TiO2 with anatase phase for the degradation of phenol with visible light (λ N 400 nm). Under the conditions tested, mineralisation of phenol was signi?cantly higher at 400 °C due to its smaller particle size and higher adsorption area toward the organic substrate. Hong et al. [56] examined the in?uence of calcination temperatures (200–700 °C) on the photocatalytic degradation of phenol using TiO2 prepared by the sol–gel method. Under the experimental conditions, the maximum degradation was observed at a calcination temperature of 400 °C with anatase phase TiO2. The degradation of phenol decreases drastically above 600 °C which was attributed to increased crystallite size and rutile phase of titania. Lv and Lu [98] studied the effect of calcination temperatures (200–800 °C) on the photocatalytic degradation of phenol using pure anatase and rutile TiO2 in aqueous suspension, in the absence and presence of NaF. Under the experimental conditions (Anatase TiO2 = 1.0 g/L, pH = 3), the maximum photocatalytic degradation rate was shown to be at 600 °C under UV irradiation (λ N 320 nm). However, the degradation rate was retarded in the presence and absence of NaF when rutile TiO2 was used. XRD analysis revealed that the rutile phase appears at calcination temperatures up to 700 °C, and almost all the anatase crystallizes to rutile at 800 °C. As the calcination temperature increased, the particle sizes of both anatase and rutile increased according to the diffraction intensity. Chiou et al. [62] examined the ef?ciency of 0.072 mol% Pr-doped TiO2 for the photocatalytic degradation of phenol over a range of calcination temperatures from 100 to 800 °C. XRD patterns indicated that the samples annealed at 100–400 °C become crystalline with dominantly anatase structure. Upon annealing at 500–600 °C, the samples possess both anatase and rutile structures, whereas the samples annealed at 800 °C are completely transformed to the rutile structure. The highest degradation is achieved with samples annealed at 600 °C (BET = 40 m2/g) compared to the samples annealed at 800 °C. There have been several investigations on photocatalytic degradation of various dyes by titania nanoparticles, however, information on the titania ?lm fabrication routes on the photocatalytic degradation of dyes is still desirable. Song et al. [99] investigated the effect of titania ?lms with different nanostructures of nanorods (NR), sol–gel ?lm (SG), nanotubes (NT) and nanoparticle aggregates (DP) on the photocatalytic degradation of rodamine B (RB), methylene blue (MB) and methyl orange (MO). In the case of RB degradation, the observed order was reported to be NR N DP N SG N NT, and the order was shown to be NR N SG N DP N NT for the degradation of MB. An order of NT N DP N SG N NR was observed for the degradation of MO. The loss of TiO2 nanoparticle surface area throughout the growth

Table 7 In?uence of doping concentration on the photocatalytic degradation of various phenols. Pollutant Phenol 4-Chlorophenol 2,4-Dichlorophenol p-chlorophenol p-chlorophenol Phenol 4-Chlorophenol 2-Chlorophenol Phenol Phenol 2,4,6-Trichlorophenol Light source Visible Visible Visible UV UV UV UV UV UV UV UV-A Photocatalyst Fe3+–TiO2 Ce–TiO2 V–TiO2 Ag–TiO2 Pd–TiO2 Fe3+–TiO2 Zr4+–TiO2 Co–TiO2 La–TiO2 Pr–TiO2 Ag–TiO2 Doping (%) 0.4–5.1 0–1.0 0.25–5.0 0.1–0.5 0.1–0.5 0–3.0 0.5–5.0 0.004–0.14 1.0–5.0 0–0.22 0–2.0% Optimum doping (%) 1.0 0.6 1.0 0.5 0.5 0.5 3.0 0.036 1.0 0.072 0.5 Ref. [102] [103] [104] [101] [101] [105] [106] [61] [89] [62] [107]

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of nanocrystallites during high temperature calcination (N400 °C) and serious aggregation of prepared nanoparticles when dispersed in aqueous solution need to be addressed to achieve further progress. 3.9. Doping and mixed semiconductor The recombination of photogenerated electrons and holes is the reason behind the low photoactivity of TiO2. Structural imperfections in the TiO2 lattice generate trap sites which may act as recombination centres, leading to a decrease in the levels of electrons and holes. In order to enhance the TiO2 photocatalysis as well as the response into the visible spectrum of solar light, TiO2 has been doped with certain transition metals, non-metals and ionic components. Doped ions can also act as charge trapping sites and thus reduce electron–hole recombination. The effect of doping on photocatalytic activity is governed by several factors, e.g. the type and concentration of dopant, preparation method, the structure and the initial concentration of the pollutants (phenols and its derivatives), and physico-chemical properties of the catalyst. Both positive and negative results have been reported from doping with metal ions. The increase in charge separation ef?ciency will enhance the formation of both free hydroxyl radicals and active oxygen species [100]. Table 7 shows the in?uence of doping on the photocatalytic degradation of various phenolic compounds. Kirilov et al. [101] studied the ef?cacy of Ag and Pd modi?ed TiO2 prepared by impregnation method for the photocatalytic degradation of p-chlorophenol. Under the conditions tested, complete degradation is achieved after 180 min reaction in the case of 0.5%Ag/TiO2 and 0.1%Ag/TiO2. For 0.25%Ag/TiO2, the complete degradation is observed after 240 min. In contrast, 90% degradation is obtained with unmodi?ed TiO2. Similar trends have also been observed for 0.1%Pd/TiO2 and 0.5%Pd/TiO2. Comparatively Ag/TiO2 samples showed higher photocatalytic activity than Pd/TiO2. In comparison to unmodi?ed TiO2, the improved activity of Ag and Pd modi?ed TiO2 is attributed to better charge separation and consequently, to slower recombination. Rengaraj and Li [107] investigated the photocatalytic activity of Ag– TiO2 prepared by the sol–gel method for the degradation and mineralisation of 2,4,6-trichlorophenol (TCP) in aqueous solution under UV-A illumination. In Ag-doped TiO2, the Ag content is varied from 0.1 to 2.0 wt.%. Degradation of TCP is demonstrated to be more than 95% within 120 min, and the corresponding TOC reduction is reported to be 80% in 120 min. The kinetics of TCP degradation follows a pseudo-?rst-order kinetic. The 0.5%Ag–TiO2 catalyst is shown to have the highest ef?ciency for the degradation of 2,4,6-TCP. The degradation rate constant is reported to be 0.1046 min? 1 for 0.5%Ag– TiO2 compared 0.0682 min? 1 for undoped TiO2. The photocatalytic activity enhancement due to Ag doping is related to better separation of photogenerated charge carriers and improved oxygen reduction inducing a higher extent of degradation of aromatics. Arabatzis et al. [108] observed a two times faster degradation of methyl orange in the presence of Au/TiO2 thin ?lms compared to that obtained with the original TiO2 material. TiO2 with gold surface coverage of 0.8 μg cm? 2 is shown to be more ef?cient for photocatalysis. The improvement in the ef?ciency is reported to be more than 100%. This enhancement is attributed to the presence of Au particles which attract conduction band electrons and prevent electron–hole recombination. Sonawane and Dongare [91] tested the effect of Au doping (1–2% Au/TiO2) on the photocatalytic degradation of phenol in solar light irradiation of 4.5–5 W m2/day. In comparison to undoped TiO2, TiO2 doped with 1–2%Au showed 2–2.3 times higher photocatalytic activity. The degradation rate constant of pure TiO2 was 0.004945 min? 1, whereas the rate constants of 1%Au/TiO2 and 2%Au/ TiO2 were 0.01041 min? 1 and 0.01128 min? 1, respectively. The increase in photocatalytic activity of Au-doped TiO2 ?lms was observed. This increase is attributed to the combined effect of shift in absorption wavelength to visible, the shift in Fermi level to more negative potentials and improved ef?ciency of the interfacial charge

transfer process. Using the sol–gel method, Venkatachalam et al. [106] examined the effect of Zr4+ doping (0.5–5.0 mol% Zr4+) onto TiO2 matrix for the photocatalytic degradation of 4-Chlorophenol. The photocatalytic activity of 3.0 mol% Zr4+-doped TiO2 samples were reported to be higher than that of nano TiO2 and Degussa P25. Enhanced adsorption of 4-CP over the catalyst surface and the decrease in particle size are indicated to be the reason for high activity of the catalyst. The entry of Zr4+ into the lattice of TiO2 creates charge compensating anion vacancies in the lattice points of TiO2 which may enhance the adsorption of 4-CP. Zr4+-doped TiO2 contains only anatase phase whereas pure TiO2 contains both anatase and rutile phase together. The incorporation of Zr4+ into TiO2 led to small grain size, large surface area and high band gap values. This also led to the formation of more electron capture traps which contribute to high separation ef?ciency of photogenerated carriers. Burns et al. [109] have studied photodegradation of 2-chlorophenol using sol–gel synthesized Nd-doped TiO2 under UV irradiation. They reported that doping TiO2 with Nd3+ reduces the degradation time because of the difference in the ionic radii of Nd3+ and Ti4+. Much larger substitutional Nd will cause localised charge perturbation and formation of oxygen vacancies which act as electron traps. Bellardita et al. [110] investigated the effect of metal loading (W, Co and Sm) onto a TiO2 matrix on the photocatalytic degradation of 4nitrophenol. The photocatalytic activity of the samples with increasing metal loading is shown to increase until a maximum value and afterwards decreases. In comparison to Degussa P25, TiO2 loaded with 1% W and Sm showed a signi?cant improvement in the photocatalytic activity under UV irradiation, whereas Co is reported to be bene?cial only under visible light irradiation. This effect was ascribed to an increased charge separation of the photogenerated electron–hole pairs in the presence of dopant ion. Sobana et al. [111] studied the effect of activated carbon (AC) content onto ZnO for the photocatalytic degradation of 4-acetylphenol in the presence of UV irradiation. Bare ZnO gives a complete disappearance of 4-AP in about 240 min of UV irradiation. In contrast, 9AC-ZnO removes 4-AP completely from the solution within 150 min due to a synergistic effect between ZnO and activated carbon. The distribution of ZnO is found to be more homogeneous in 9AC-ZnO compared to other AC-ZnO catalysts. 9AC-ZnO is found to be more ef?cient in solar light than in UV light. Wang et al. [112] studied the Ag-doped ZnO for the photocatalytic degradation of MB by varying the doping ratio of Ag/Zn from 0 to 2%. Under the conditions tested, Ag doping signi?cantly improved the photocatalytic activity of ZnO due to O vacancies, crystal de?ciencies and increased speci?c surface area. Barakat et al. [61] studied the effect of Co (0.004–0.14 mol%) doping onto TiO2 nanoparticles by a sol–gel technique from TiCl4 for the photocatalytic degradation of 2Chlorophenol. The photocatalytic degradation ef?ciency of 0.036 mol % Co-doped TiO2 was shown to be 75.4% higher compared to undoped TiO2 under the conditions tested. A further increase in Co concentration to 0.14 mol% resulted in a decrease in the degradation ef?ciency to 83%.They reported that catalyst samples with 0.036 mol% Co particles annealed up to 400 °C were amorphous, and became crystalline with dominantly anatase structure upon annealing at 600 °C. Under irradiation, Co(III) ions in the Co-doped TiO2 work as electron scavengers which may react with the superoxide species and prevent the hole–electron recombinations and consequently increase the photo degradation ef?ciency. Rengifo-Herrera and Pulgarin [113] tested the photocatalytic activity of N,S co-doped and N-doped commercial anatase TiO2 powders for the degradation of phenol under simulated solar irradiation. Undoped P25 was shown to have the highest photocatalytic activity. N,S co-doped powders showed almost the same photocatalytic activity as undoped TKP 102 (Tayca) while N-doped TKP 102 was the less active photocatalyst due to N impurities on the TiO2 acting recombination centres. Chiou et al. [62] studied the effect of Pr3+ dopant concentration (0.018–0.22 mol%) on the photocatalytic degradation of phenol. Using the undoped sample,

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42% of 50 mg/l phenol was degraded after 2 h irradiation. The degradation ef?ciency increases with increasing Pr concentration. The highest degradation is observed to be 94.4% with the sample containing 0.072 mol% Pr3+ that was annealed at 600 °C. A further increase in Pr concentration to 0.22 mol% leads to a slight decrease of degradation ef?ciency to 91.3%. In the Pr-doped TiO2 particles, Pr+ 3 works as an electron scavenger; this may react with superoxide species and prevent the hole–electron recombinations, and thus increase photooxidation ef?ciency. However, the degradation was lower than 80% with undoped and other metal-doped TiO2 particles under similar conditions. Zaleska et al. [114] studied the photocatalytic activity of TiO2 modi?ed with boric acid for the degradation of phenol under UV and visible light respectively. The use of boric acid as a boron source resulted in lower photoactivity of the obtained BTiO2 than pure TiO2 under UV light. All tested TiO2 powders with boric acid series have almost the same photoactivity under visible light. Naeem and Ouyang [105] studied the effect of Fe3+ doping (0–3.0 mol% Fe3+) on the photocatalytic activity of TiO2 for the degradation of phenol under UV light. Fe3+-doped TiO2 was reported to possess the anatase structure with a range of crystal size of 8–11 nm. The highest degradation ef?ciency was found at 0.5 mol% Fe3+-doped TiO2. Liqiang et al. [89] tested the ef?cacy of La-doped TiO2 prepared by sol–gel method for the photocatalytic degradation of phenol. After 2 h calcination at 600 °C, the order of photocatalytic activity of La-doped TiO2 for phenol degradation was demonstrated to be 1 N 1.5 N 3 N 0.5 N 5 N 0 mol% La. This is related to the La dopant which could inhibit the crystallite growth and the phase transformation of TiO2 nanoparticles so that the surface content of oxygen vacancies or defects increased. Venkatachalam et al. [53] reported that, under identical and optimal experimental conditions, 1 mol% Mg2+ and Ba2+ doped nano TiO2, Nano TiO2 and P-25 required 300, 360 and 450 min respectively for complete mineralisation of 4-Chlorophenol. The photocatalytic degradation of 4-CP over TiO2 and Mg2+ and Ba2+ doped TiO2 indicated the higher activity for doped TiO2. Enhanced adsorption of 4-CP on the catalyst surface and smaller particle size due to Mg2+ and Ba2+ loadings are indicated to be the cause for higher activity of the catalysts. Lee et al. [115] studied the photocatalytic activity of carbon-doped TiO2 powders synthesized by the sol–gel method for the degradation of phenol under visible light (λ N 420 nm) irradiation. The particles contain a certain fraction of oxygen vacancies in the anatase structure as evidenced by XPS results. Under the conditions tested, carbon doped TiO2 was reported to be ef?cient for the degradation of phenol. The high photocatalytic activity of C-doped TiO2 was attributed to the band gap narrowing effect by vacancy states in the anatase structure stabilized by carbon doping. Doping shifted the absorption edge to a lower energy level, thus effectively increasing the photocatalytic activity in the visible–light region. Tian et al. [104] studied the photocatalytic activity of V-doped TiO2 for the degradation of MB and 2,4-dichlorophenol using UV and visible light. In the presence of UV light, 0.5% V-TiO2 was reported to be ef?cient for the degradation of MB relative to pure TiO2. Under visible light irradiation, the degradation rate of 2,4-dichlorophenol over 1% V-TiO2 is shown to be two times higher compared to undoped TiO2. Yuan et al. [73] studied the in?uence of co-doping of Zn2+ and Fe3+ on the photocatalytic degradation of phenol under solar light irradiation. The co-doping of 0.5 mol% Zn2+ and 1 mol% Fe3+ onto TiO2 was found to be two times more ef?cient compared to pure TiO2. This behaviour was attributed to the coupled in?uence of the co-dopant and titania energy bands. Lee et al. [116] compared the photocatalytic degradation of p-nitrophenol over TiO2 and TiO2/SiO2 nanoparticles prepared by the micro-emulsion method using PFPE-NH4 surfactant. TiO2/SiO2 (80:20) nanoparticles were reported to have a higher photocatalytic activity than pure TiO2 and the TiO2/SiO2 (90:10) particles. This effect has been attributed to the decrease in crystallite size from 17 to 13 nm with an increase in silica content. Under the conditions tested, Silva

et al. [103] noticed an optimum Ce content of 0.6% (w/w) into the TiO2 matrix for enhanced degradation of 4-chlorophenol in the visible light. In addition to the retardation of phase transformation, this behaviour was related to the shift of TiO2 absorption edge towards longer wavelengths, by reducing the band gap of the original material. Wang et al. [117] compared the photocatalytic degradation of phenol in the presence of N2-doped TiO2 and P25 under visible and sunlight irradiation. The photocatalytic activity of the N2-doped TiO2 with anatase phase is shown to be higher than that of the P25 under visible light irradiation. P25 exhibits higher photoactivity for the degradation of phenol compared to N2 doped TiO2. Kim et al. [118] studied the photocatalytic degradation of 4-CP using anatase, rutile, Ni 8 wt.%-doped TiO2 powders (anatase and rutile) under UV and visible light irradiation. TEM and EDS analysis indicated that the added Ni atoms were distributed in the rutile TiO2 lattice which had an average grain size of 2–4 nm. BET analysis showed that Ni-doped powder has a larger surface area (234 m2/g) than the other powders (100–150 m2/ g).The UV/VIS-DRS absorption showed that the nano-sized Ni-doped powder had a higher wavelength range (480–500 nm) than the anatase and rutile powder (380–500 nm). Under both UV and visible light irradiation, the effectiveness of Ni 8 wt.%-doped TiO2 (234) for the photocatalytic degradation of 4-CP was much higher than that of other powders (100–150 m2/g). Adan et al. [102] investigated the effect of Fe3+ doping (0.4–5.1 wt.%) on the photocatalytic degradation of phenol. The photocatalytic activity is shown to be enhanced upon doping with Fe3+ up to ca. 1 wt.%. Doping above ca. 3 wt.% does not produce any enhancement of the catalytic activity. Nahar et al. [119] examined the dependence of the photocatalytic degradation of phenol on Fe content under UV and visible light irradiation. An increase in the Fe3+ content (Fe/Ti, x = 0.002 to 0.005) is reported to increase the photocatalytic activity, however, the overloaded Fe (x = 0.008, 0.01) considerably decreased the rate, even below the undoped TiO2. This effect was ascribed to the balance of excited electron/hole trapped by the doped Fe3+ and their charges recombination on the doped Fe3+ level. The molar ratio of Fe/Ti (x = 0.005) was optimum for both the UV and solar light irradiations. The UV activity was 4.3 times higher than that of the visible light activity. Gorska et al. [120] compared the ef?ciency of phenol degradation for N,C doped-TiO2 prepared by TIP hydrolysis under visible light to P-25 irradiated by UV light. In the Vis/ N,C–TiO2 and UV/P-25 systems, nearly 60 min irradiation resulted in 69 and 76% of phenol degradation respectively. However, photocatalytic activity under visible light (N400 nm) of N,C–TiO2 was four times higher than photoactivity of P-25 TiO2. Phenol degradation ef?ciency for commercial TiO2/UV and doped TiO2/Vis systems was reported to be comparable. 4. Conclusion Based on recent representative investigations, the role of various operating parameters on the photocatalytic degradation of various phenols and substituted phenols has been explored in this review. TiO2 has been suggested to be ef?cient for the degradation and mineralisation of various toxic organic pollutants such as phenols and dyes in wastewater water in the presence of UV, visible or solar light and oxygen. The ?ndings also suggest that various operating parameters such as photocatalyst type, light intensity, pollutant type and initial concentration, catalyst amount, initial pH of the reaction medium, catalyst application mode, oxidizing agents/electron acceptors, and the presence of ionic components in solution can in?uence signi?cantly the photocatalytic degradation rate of phenols. Optimising the degradation parameters is crucial from the perspective of ef?cient design and the application of photocatalytic oxidation processes to ensure sustainable operation. The application of this technique on actual mixed matrices merits further investigation to yield stable pollutant removal through the optimisation of process parameters. Metal and non-metal doped TiO2 have been reported to

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result in improved degradation rates. While doped TiO2 has shown signi?cant improvement in the ef?ciency of TiO2 to remediate wastewater in laboratory scale systems, it is evident that the need to develop pilot scale treatment systems and to apply the technique in cost effective wastewater puri?cation processes requires continued in-depth research. In spite of extensive investigations, the commercial exploitation of photocatalysis has been hindered by the lack of ef?cient and low cost visible light harvesting catalysts, a relatively poor understanding of reactor design criteria, and inadequate scale up strategies. Future research should focus on the development of a more reliable photocatalyst that can be activated by visible and solar light or both. In addition, more work is required on the modelling of the photoreactor to optimise its design for pollutant degradation. In the literature, there is currently little information on this aspect. Present research activities at CQ University and QUT, Australia focus on the computational ?uid dynamics modelling of a ?at plate reactor to optimise its design and to predict its performance. Although this review is not comprehensive in the scope of the photocatalytic degradation of organic pollutants, it does however address the fundamental principles and recent applications in this area. Acknowledgement This study is supported under an Australian Research Council (ARC) linkage grant in collaboration with CM Concrete Private Limited and Department of Public Works, QLD Government. The authors gratefully acknowledge the ?nancial support of the ARC project. One author is also grateful for the ?nancial support of the Queensland Government through the Smart State fellowship scheme. References
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