Exposisicon Martes-13 - Marc Pera-AOPs

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Applied Catalysis B: Environmental 47 (2004) 219256

Degradation of chlorophenols by means of advancedoxidation processes: a general review

Marc Pera-Titus a, Vernica Garca-Molina a, Miguel A. Baos b,1,Jaime Gimnez a, Santiago Esplugas a,

a Department of Chemical Engineering and Metallurgy, Faculty of Chemistry, University of Barcelona, Mart i Franqus 1, 08028 Barcelona, Spainb Department of Physical Chemistry, Faculty of Chemistry, University of Barcelona, Mart i Franqus 1, 08028 Barcelona, Spain

Received 20 April 2003; received in revised form 25 July 2003; accepted 9 September 2003

Abstract

Advanced oxidation processes (AOPs) constitute a promising technology for the treatment of wastewaters containing non-easily removableorganic compounds. Chlorophenols (CPs) are a group of special interest due to their high toxicity and low biodegradability. Data concerningthe degradation of CPs by means of AOPs reported during the period 19952002 are evaluated in this work. Among the AOPs, the followingtechniques are studied: processes based on hydrogen peroxide (H2O2 + UV, Fenton, photo-Fenton and Fenton-like processes), photolysis,photocatalysis and processes based on ozone (O3, O3 + UV and O3 + catalyst). Half-life times and kinetic constants for CP degradation arereviewed and the different mechanistic degradation pathways are taken into account. 2003 Elsevier B.V. All rights reserved.

Keywords: Chlorophenol; AOP; Photolysis; Fenton; Photo-Fenton; Ozone; Photocatalysis; Radiation; Pathway; Half-life time; Kinetic constant

1. Introduction

In the beginning of the 21st century, the mankind has toface the problem of water as an important threat. Accordingto WHO [135], the shortage or even lack of water affectsmore than 40% of the world population due to political,economical and climatological reasons. Besides, more than25% of the world population suffers from health and hy-gienic problems related to water. Despite the plans carriedout by UNO in recent years, 1100 million people have stillno access to improved water supply and sanitation, espe-

cially concentrated in underdeveloped countries of Africa,Asia and Latin America [248].On the other hand, the domestic use and industrial ac-

tivity, of especially impact among the developed countries,generate high amounts of residual wastewater, whose directdisposal to natural channels causes a considerable effect inthe environment. This fact, together with the need to re-store this water for new uses, makes practically essential the

Corresponding author. Tel.: +34-93-4021288; fax: +34-93-4021291.E-mail address: [email protected] (S. Esplugas).1 Tel.: +34-93-4021220; fax: +34-93-4021231.

purification of wastewater to achieve the desired degree ofquality.

Because of an increasing social and political concern onenvironment, the research field of water purification has beenextensively growing in the last decades, comprising bothpolluted wastewaters and groundwaters from seas, rivers andlakes, as water quality control and regulations against haz-ardous pollutants have become stricter in many countries.More recently, reflecting a new environmental conscience,the European Directive 2000/60/CE [60] stresses the need toadopt measures against water pollution in order to achieve

a progressive reduction of pollutants.On this way, chlorophenols2 (CPs) constitute a particulargroup of priority toxic pollutants listed by the US EPA in theClean Water Act [67,102,131] and by the European Deci-sion 2455/2001/EC [57], because most of them are toxic andhardly biodegradable, and are difficult to remove from theenvironmentthe half-life in water can reach 3.5 months inaerobic waters for PCP and some years in organic sediments

2 As generally submitted in the literature, the general nomenclatureused for CPs is the following one: monochlorophenol (MCP), dichlorophe-nol (DCP), trichlorophenol (TCP), tetrachlorophenol (TTCP) and pen-tachlorophenol (PCP).

0926-3373/$ see front matter 2003 Elsevier B.V. All rights reserved.doi:10.1016/j.apcatb.2003.09.010


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220 M. Pera-Titus et al. / Applied Catalysis B: Environmental 47 (2004) 219256

[1]. Because of their numerous origins, they can be foundin ground waters, wastewaters and soils [244] and even inthe trophic chain of places with very low pollution levels[193,247]. As pointed out in Table 1, they might producedisagreeable taste and odor to drinking water at concentra-tions below 0.1g l1 [238] and adverse effects on the en-

vironment [79].All CPs possess bactericidal activities, phytotoxicity andability to bioaccumulate in organisms that increase with in-creasing the chlorination and substitution away from theortho-(2-) position (see Table 1). The higher toxicity of themore chlorinated CPs may be ascribed to an increase in ly-pophility which leads to a greater potential for uptake intothe organism. The ortho-substituted congeners are gener-ally of lower toxicity than the meta- and para-ones, becauseortho-substituted chlorine seems to shield OH group, whichapparently interacts with the active site in aquatic organ-isms [92]. On the other hand, toxicity also depends on theextent to which CP molecules are dissociated, with an in-

creasing toxicity when the pH decreases, because the moretoxic non-dissociated forms predominate at lower pH.

The presence of CPs has been detected in both surface andground waters [114]. Toxic referencevalues of 43.8, 36.5 and13.0g l1 are, respectively, suggested for 2-MCP, 2,4-DCPand PCP in surface waters, and maximum average valuesshould not exceed 2.020, 4.380 and 0.055 mg l1 [68,125].The limiting permissible concentration of CPs in drinkingwater should not exceed 10g l1 [220].

Table 1Literature guidelines for CPs in drinking water [6971] and toxicity based on Daphnia magnaa [58] and Rat [70,235]

CP congeners Guidelines (g l1) Daphnia magna 24 h-LC50b (mg l1) Rat 24 h-LC50c (mgkg1)

Range Mean Oral Subcutaneous

2-MCP 0.10 16.6019.30 17.95 670 9503-MCP 0.10 13.8017.79 15.78 570 10304-MCP 0.04 5.7910.34 8.07 261 1390

2,3-DCP 0.101.00 4.096.30 5.19 2,4-DCP 0.50 2.482.89 2.68 580 17302,5-DCP 0.20 4.50 2,6-DCP 0.30 8.6910.08 9.38 2940 17303,4-DCP 1.00 2.552.98 2.77 3,5-DCP 1.852.33 2.09

2,3,4-TCP 2.002.48 2.24

2,3,5-TCP 2.062.50 2.28 2,3,6-TCP 6.258.52 7.38 2,4,5-TCP 0.122.00 1.882.29 2.08 820 22602,4,6-TCP 4.936.01 5.47 2800 3,4,5-TCP 0.820.93 0.88

2,3,4,5-TTCP 1.521.98 1.76 572 2,3,4,6-TTCP 1.00 2.70 140 2102,3,5,6-TTCP 1.872.66 2.27 109

PCP 30.0 0.620.89 0.76 50 100

a The Daphnia magna were all >72h old. The water hardness was 200mgl1 as CaCO3, dissolved oxygen >2.27 mg l1, pH 7.88.2 and temperature20 C.

b At this concentration of CP, 50% of Daphnia magna die after 24 h contact with pollutant.c Estimated oral 24 h-LD50 for human beings is between 50 and 500 mg kg1.

CPs are introduced into the environment as a result of sev-eral man-made activities. Because of their broad-spectrumantimicrobial properties, CPs have been used as preservativeagents for wood, paints, vegetable fibers and leather and asdisinfectants. In addition, they have been widely employedin many industrial processes as synthesis intermediates or as

raw materials in the manufacturing of herbicides, fungicides,pesticides, insecticides, pharmaceuticals and dyes. CPs maybe also generated as by-products during waste incineration,the bleaching of pulp with chlorine, and in the dechlorina-tion of drinking water [259].

According to [235], the world market for CPs is fairly sta-ble and is ca. 100 kt per year. The production of heavy andlight CPs is ca. 2530 and 60 kt per year, respectively. M-, D-and TCPs which have no chlorine atoms in meta position rel-ative to OH are industrially produced by direct chlorinationof phenol with chlorine gas; CPs having at least one chlorineatom in the meta position must be produced by other typesof reactions, such as hydrolysis, sulfonation, hydrodechlori-

nation, hydroxylation, and alkylation; and TTCPs and PCPare produced batchwise by means of the chlorination of lesschlorinated CPs in the presence of aluminum or iron trichlo-ride. Emissions are mainly due to the manufacture, storage,transportation and application of CPs.

The need to restore contaminated sites to avoid furtherrisks to the environment has aroused in the last few yearsthe development of effective methods for CP removal. Themain goal is to achieve a complete mineralization to CO2


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and H2O in addition to smaller amounts of some ions, e.g.chloride anions, or at least to produce less harmful interme-diates. The conventional pollutant destructive technologiesinclude biological, thermal and chemical treatments [124].The former usually require a long residence time for mi-croorganisms to degrade the pollutant, because they are af-

fected by CP toxicity; thermal treatments present consider-able emission of other hazardous compounds; and the latter,which include processes as flocculation, precipitation, ad-sorption on granular activated carbon (GAC), air stripping orreverse osmosis (RO), require a post-treatment to remove thepollutant from the newly contaminated environment [54].

Alternative methods to these well-established techniquesinvolve the oxidation of CPs with reagents such as airor oxygen in wet oxidation and supercritical wet oxida-tion [146,149,152], electrons in electrochemical oxidation[52,117,205], potassium permanganate, chlorine, hydrogenperoxide and ozone [42,256]. Among these techniques,the so-called advanced oxidation processes (AOPs) [86,89]

appear to be a promising field of study, which have beenreported to be effective for the near ambient degradationof soluble organic contaminants from waters and soils,because they can provide an almost total degradation[6,25,26,39,40,86,95,111,112,147,171,179,181,195,235].Even though these techniques can provide the conver-sion of contaminants to less harmful compounds, usuallyoxygenated organic products and low molecular acids[83,104,145], they are limited to treat waters which con-tain low concentrations of organic or inorganic scavengingmaterial [88]. Experiences with different oxidation tech-nologies and substrates have shown that a partial oxidation

of toxic water may increase its biodegradability up to highlevels [133,214].Highly reactive hydroxyl radicals (HO) are traditionally

thought to be the active species responsible for the destruc-tion of pollutants [34,86,87,97,195]. Thanks to its high stan-dard reduction potential of 2.8 V vs. NHE3 in acidic media(see Table 2) [77,245], these radicals would be able to ox-idize almost all organic compounds to carbon dioxide andwater, except for some of the simplest organic compounds,such as acetic, maleic and oxalic acids, acetone or simplechloride derivatives as chloroform [28], which are of a veryinteresting kind because they are typical oxidation productsof larger molecules after fragmentation and they take part inenergetic cycles of most living organisms. They should beextremely unstable and continuously generated by chemi-cal, photochemical or electrochemical reactions. Dependingon the nature of organic species, two types of initial attackmight be possible: it might abstract a hydrogen atom in thecase of alkanes and alcohols, or it might attach itself to amolecule in the case of aromatic compounds, such as CPs.

Even though AOPs for water and wastewater treatmentshow high efficiencies, they actually work at high cost

3 Normal standard hydrogen electrode.

Table 2Standard reduction potentials of some oxidants in acidic media [120]

Oxidant Standard reductionpotential (V vs. NHE)

Fluorine (F2) 3.03Hydroxyl radical (HO) 2.80Atomic oxygen 2.42

Ozone (O3) 2.07Hydrogen peroxide (H2O2) 1.77Potassium permanganate (KmnO4) 1.67Hypobromous acid (HbrO) 1.59Chlorine dioxide (ClO2) 1.50Hypochlorous acid (HClO) 1.49Chlorine (Cl2) 1.36Bromine (Br2) 1.09

[190]. They only appear to be suitable for COD contentslower than 5gl1, since higher COD contents would re-quire the consumption of too large amounts of reactantsinthose cases, it would be more convenient to use wet oxida-tion or incineration [184]. Furthermore, the combination ofan AOP as a preliminary treatment, followed by an inex-pensive biological process, seems to be an interesting op-tion from an economical point of view [129,169,213,254].A scheme of this kind of processes is depicted inFig. 1.

This work is devoted to summarize the main aspects deal-ing with their degradation and mineralization of CPs bymeans of AOPs. It is specially focused on the collection ofhalf-life times and kinetic constants for the processes consid-

ered, which appear in the literature in the period 19952002,and on the comparison between the different mechanisticpathways concerning CP degradation. Among the variousAOPs proposed in the literature, which even include tech-niques based on ultrasound (Lin et al., 2000), plasma [231],or electrohydraulic discharge [249], the main ones takeninto account in this work are based on the following ones[89,199]:

(1) Photolysis (UV or VUV).(2) Hydrogen peroxide (H2O2):

H2O2 + UV;

Fenton: H2O2 + Fe2+/Fe3+; Fenton-like reagents: H2O2+Fe2+-solid/Fe3+-solid; photo-Fenton: H2O2 + Fe2+/Fe3+ + UV.

(3) Ozone (O3): ozonation: O3; photo-ozonation: O3 +UV; ozonation + catalysis: O3 + H2O2 and O3 +

Fe2+/Fe3+.(4) Heterogeneous catalysis+ UV and photocatalysis.

Heterogeneous catalysts and photocatalysts reviewed:TiO2 + CdS+ combinations.


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Fig. 1. Scheme of a process which combines an AOP with a biological process for wastewater treatment [190].

2. Analytical topics

Different analytical methods based on particular ana-lytical techniques have been proposed for qualitative andquantitative analysis of CPs and reagents concerning theirdegradation. These methods usually involve aeration or de-saeration, extraction of the aqueous sample with an organicsolvent and filtration in case a heterogeneous catalyst orsolid reactant is employed. The most popular techniques

for qualitative and quantitative analysis of CPs and the or-ganic and inorganic species involved in their degradationby AOPs are summarized in Table 3.

3. Reactor and lamp configuration

3.1. Reactor configuration for direct photolysis and AOPs

based on hydrogen peroxide

As reported in most of the papers reviewed [21,22,143,166,220,257], completely mixed batch cylindrical glassvessels constitute the most popular reactor configurationfor the studies of degradation of CPs for Fenton studies.Other batch configurations concern erlenmeyers or cylin-drical glass flasks [17,119,161,208] (Lu et al., 2002) orglass bottles [163]. As shown in Fig. 2a, a typical batchoperation would consist in a chemical storage and dosing

Fig. 2. Flow chart for (a) the Fenton process and (b) the photo-Fenton process [17,22].

modules (for H2O2, FeSO4, acid, and lime/NaOH) and areactor, provided with the required elements to allow theintroduction of probes and reagents, and usually surroundedby a cooling jacket. The materials of construction for thereactor and holding tank are typically types 304 or 316stainless steel, while those for the chemical storage systemsmay also be HDPE [242].

On the other hand, tubular configurations are usually em-ployed for direct photolysis and photo-Fenton processes, and

processes based on H2O2/UV reagent, in order to achieve agood interaction between CPs and other intermediates andradiation [17,220]. All experimental measures must be takenfrom treated stream, due to the concentration profiles whichappear inside the reactor. A scheme for the photo-Fentonprocess is shown in Fig. 2b.

3.2. Reactor configuration for AOPs based on ozone

According to the literature, semi-batch stirred reactorsglass bottles or cylindrical vesselsare typical reactor con-figurations for ozonation processes [1,2224,53,140,141].On the other hand, bubble columns, wetted-sphere adsor-bers, and other special designs have also been employed[55,100,186,233,239] to improve the contact between thegaseous ozone stream and the liquid and polluted stream.An ozone generator must be located near the reactor to sup-ply continuously the ozone required for the reactionozone


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Table 3Analytical techniques employed in degradation of CPs with AOPs

Compound Analytical techniques References

CPs and intermediates GC [7,45,55,59,125,154,173,177,204,253,255,257]GC/MS [44,103,119,124,143,239]HPLC [1,7,2123,43,46,61,93,100,105,106,124,156,172,188,210,233]

Ion chromatography [7]Spectroscopy [220,255] (Hautaniemi, 1998)TOC analysis [1,17,24,46,61,93,105,124,139,163,166,203,230,260] (Chen and

Chan, 1996)

Cl Ion selective electrode [100,140,166]Colorimetry [141]Potentiometric titration [45,260]Capillary electrophoresis [119]Ion chromatography [1,105,139,174,233]

H2O2 Spectrophotometry [51,100,233]Iodometry [22,166]Photometry [10]Colorimetry [119]DPD method Lu et al. (2002)

Fe2+/Fe3+ Spectrophotometry [17,163,166,208]Iodometry [119,257] (Lu et al., 2002)

O2 DO meter [43,143]Oxygen electrode [61,166]

O3 Gas:Spectrophotometry [99,100,233]Iodometry [23,204]Colorimetry [24,170]Electro-magnetically [188]

DissolvedIndigo method [11,12,23,44,55,100,188,233]Titration method Kuo (1999)Iodometry [24]

Cations Cationic chromatography [17]pHa pH-meter

Metals AAS [163]

TiO2 XRD [59]

Light intensity Actiometry [100,233,257]Radiometry [59]

a pH change during degradation processes are often monitored to ensure that they are progressing as planned.

cannot be stored due to the potential flammable character ofozoneoxygen mixtures. The gaseous ozone stream is usu-ally introduced into the reactor through a porous glass-platediffuser, where it dissolves, and the residual one in the out-flow stream must be removed. In addition, a lamp or severalones must be added in the contact system to allow matterinteract with both ozone and radiation, a scheme which isshown in Fig. 3.

Ozone can be generated from pre-treated air or oxygen byirradiating with UV or by means of an electrical discharge[65,235]. In both cases, high energy is required and the ef-ficiency of ozone generation is lowlower than 0.25 wt.%for the first technique, with a top flow rate of 0.5 g/h [14].The concentration in the ozonated gas is low, usually be-tween 1 and 5 wt.%. The first technique is the most ex-

tended for ozone generation in the most recent references[141,142,188].

3.3. Reactor configuration for AOPs based on

photocatalysis

Some laboratory and pilot plant photo-reactor configu-rations have been proposed for photocatalysis. The mosttypical laboratory-scale configurations include vigorouslystirred batch photochemical cells (see Fig. 4), cylindri-cal flasks or vessels, glass dishes, and annular differentialphoto-reactors [46,61,85,99,106,124,125,139,172,191,203,208,228,255].

On the other hand, pilot plant configurations may be basedon a continuous rotating disk photo-catalytic reactor (RDPR)


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Fig. 3. Flow chart for an ozonation process together with or without radiation [17,22].

[59], which are set up with a semicircular shape vessel withthe water to be treated and the catalytic pellets are loadedon the surface of the rotating disk by using a UV resistantadhesive. Moreover, these configurations may also be basedon several continuous compound parabolic collectors (CPC)at Plataforma Solar de Almeria (PSA), connected in series,

Fig. 4. (a) Scheme of a photocatalytic reactor (adapted from Chen et al. [46]). Schematic view of PSA detoxification loop (b) with CPCs modules and(c) with flat reactor [84].

with a total reflective surface of 9 m2 and 37 of inclina-tion [93,106,203] and on cylindrical tanks (flat reactor ofPSA) with several tubes located at the bottom of the reac-tor, through which circulates an air-flow [84,165]. Fig. 4band c shows the schemes of the two reactor configurationsat PSA.


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3.4. Lamp configuration: general topics

Different sorts of lamps are employed for the gen-eration of radiation supplied to CP samples for directUV-photolysis and for techniques based on UV/H2O2,UV/O3, photo-Fentons reagents and photocatalysis. Among

the various commercial radiation sources, high-, medium-and low-pressure mercury vapor lamp for the generation ofUV radiation [17,21,22,166,257] and solar-simulated xenonlamps as a source of visible radiation [208,220] are widelyemployed. Among the different configurations, the lampcan be located in an axial position housed by a sleeve [21]or in its center in a vertical manner [257].

4. UV-photolysis

Direct photolysis involves the interaction of light withmoleculesin addition to waterto bring about their disso-

ciation into fragments, with the following mechanistic path-ways [34]:

CP degradation

CP+ h Intermediates (1)

Intermediates+ h CO2 + H2O+ Cl (2)

Earlier studies on direct UV-photolysis illustrated rapiddegradation of pollutants, such as CPs, in dilute aqueoussolutions [33,252], which were subsequently confirmed byfurther studies [19,90,123,128,147,219]. On the other hand,direct photolysis has been reported to lead primarily to in-ert non-chlorinated benzenoid compounds [261]. However,

it appears to be less effective than other processes whereradiation is combined with hydrogen peroxide or ozone, orwhere homogeneous, heterogeneous catalysis or photocatal-ysis are employed.

As shown in Table 4, half-life times and pseudo-first ki-netic constants change depending on the initial concentrationof CP, pH, temperature and radiation intensity.4 As initialCP concentration increases and the other variables remainconstant, the half-life times decrease, namely the degrada-tion of CP improves. Moreover, as pH increases for constantvalues of the other variables, half-life times tend to decreaseand kinetic constants referred to CP degradation tend to in-

crease, but the kinetic constants referred to TOC degradationtend to decrease. Finally, even though the effect of radiationintensity cannot be deduced directly form data in Table 4,an increase in CP degradation can be proposed with an in-crease in radiation intensity.

On the other hand, as reported in Table 4, the degrada-tion of different CPs is faster when decreasing the number

4 Flux of photons per volume unit is measured in einsteinl1 s1,where 1 einstein = 1 mol of photons of a particular wavelength. Thevalues given in Table 5 and in the other ones in this work are mean valuesfor the wavelength ranges supplied by the lamps employed in each studyreviewed.

of chlorine atoms in alkali medium at constant values of ini-tial CP concentration, radiation intensity and temperature;the values of half-life times are between 1.1 and 30.6 mindealing with the results of Bentez et al. [23]. The resultsseem to be confirmed by the series indicated in Fig. 5, wherehigher kinetic constants for less chlorinated CPs at alkaline

pH can be observed. Otherwise, this trend seems not to beconfirmed in acidic conditions, where the kinetic constant for2,4,6-TCP degradation is higher than the one for 2,4-DCP.

The trend mentioned for the degradation of CPs is alsoobserved for the evolution of quantum yield respect to thechlorination of the aromatic ring of different CPs, as indi-cated in Fig. 6. On this way, as the chlorination of the aro-matic ring increases, a lower amount of radiation intensityis employed in its degradation to less harmful compounds orto a complete mineralization. The quantum yield for thesespecies also tends to increase with temperature and to be re-duced with the pH, because of a lower interaction of dissoci-ated species with the radiation instead of the non-dissociated

ones. An equation obtained by Bentez et al. [24], whichcan be applied for 2,4,6-TCP between 10 and 40 C is thefollowing one:

= 5.85 exp

1521

T

(mol einstein1, T in K, = 185436 nm) (3)

5. AOPs based on hydrogen peroxide

Hydrogen peroxide is a safe, efficient and easy to use

chemical oxidant suitable for wide usage on contaminationprevention. Discovered by Thenard in 1818, it was first usedto reduce odor in wastewater treatment plants, and from thenon, it became widely employed in wastewater treatment [66].However, since hydrogen peroxide itself is not an excellentoxidant for many organic pollutants, it must be combinedwith UV light, salts of particular metals or ozone to producethe desired degradation results.

5.1. Fentons reagent (H2O2/Fe2+)

The Fenton reaction is a widely used and studied catalyticprocess based on an electron transfer between H2O2 and ametal acting as a homogeneous catalyst [163,209]. By far,the most common of these ones is iron [28,242]. The reac-tivity of this system was first observed in 1894 by its inven-tor Fenton [76], but its utility was not recognized until the1930s once a mechanism based on hydroxyl radicals (HO)was proposed [200,242]. Fentons reagent can be employedto treat a variety of industrial wastes containing a range oforganic compounds like phenols, formaldehyde, pesticides,wood preservatives, plastic additives, and rubber chemicals[16,29,64,77,78,80,156159,180,185,187,196,221]. Theprocess may be applied to wastewaters, sludges, and con-taminated soils [160,162,136,180,234,240] with a reduction


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Table 4Half-life times and pseudo-first kinetic constants for degradation of CPs by direct UV-photolysis for different initial concentrations of CP

CP congener [CP]0 (mM) pH T (C) Rad. intensity(einstein l1 s1)

(nm) CP t1/2 (min) kCP (min1) Reference

2-MCP 0.78 6.0 Room 5.97 105 320400 12.0 1.52 103 a [140]2-MCP 0.78 9.0 Room 5.97 105 320400 8.5 9.00 104 a [140]2-MCP 0.40 2.5 Room 6.45 107 250 25.0 8.76 103 [233]

2-MCP 0.40 9.5 Room 6.45 107 250 4.04 102 [233]4-MCP 0.30 2.0 Room 4.80 104 185436 1.1 0.564 [23]4-MCP 0.30 2.0 25 4.80 104 185436 1.2 0.539 [21]4-MCP 0.30 9.0 25 3.52 105 185436 1.2 0.432 [21]4-MCP 0.40 2.5 Room 6.45 107 250 27.3 2.52 102 [233]4-MCP 0.40 9.5 Room 6.45 107 250 13.4 5.08 102 [233]4-MCP 0.78 6.0 Room 5.97 105 320400 8.6 9.32 103 a [140]4-MCP 0.78 9.0 Room 5.97 105 320400 7.4 5.10 104 a [140]4-MCP 0.5b 7.0 20 9.74 105 172 48.0, 200c 0.832 [98]

2,4-DCP 0.30 2.0 Room 4.80 104 185436 17.5 3.80 102 [23]2,4-DCP 0.30 2.0 25 3.52 105 185436 3.80 102 [21]2,4-DCP 0.30 9.0 25 3.52 105 185436 0.173 [21]2,4-DCP 0.78 6.0 Room 5.97 105 320400 10.3 5.90 104 a [140]2,4-DCP 0.78 9.0 Room 5.97 105 320400 6.2.0 7.00 104 a [140]

2,4-DCP 0.40 2.5 Room 6.45

107

250 70.0 6.95

103

[233]2,4-DCP 0.40 9.5 Room 6.45 107 250 14.7 4.98 102 [233]2,4-DCP 0.49 Room 2.04 104 20.0 0.130 [19]

2,3,5-TCP 0.04 7.0 Room 8.08 105 Not reached [239]2,3,6-TCP 0.12 7.0 Room 8.08 105 35.0 [239]2,4,6-TCP 0.24 7.0 Room 8.08 105 22.0 [239]2,4,6-TCP 0.30 2.0 Room 4.80 104 185436 25.2 2.60 102 [23]2,4,6-TCP 0.30 2.0 25 3.52 105 185436 2.60 102 [21]2,4,6-TCP 0.40 2.5 Room 6.45 107 250 63.0 9.48 103 [233]2,4,6-TCP 0.40 9.5 Room 6.45 107 250 34.3 2.03 102 [233]

2,3,4,6-TTCP 0.30 2.0 Room 4.80 104 185436 30.6 2.10 102 [23]

PCP 0.04 2.5 Room 6.45 107 250 8.16 102 [233]PCP 0.04 9.5 Room 6.45 107 250 5.39 102 [233]

a Rate constants referred to TOC.b This CP concentration corresponds to TOC = 36mgl1.c Half-times referred to TOC.

of toxicity, an improvement of biodegradability, and odorand color removal.

Some initial works have been issued about the degrada-tion of CPs in ground waters, wastewaters and soils. These

Fig. 5. Evolution of ln[CP] with respect to time for 4-MCP, 2,4-DCP and 2,4,6-TCP for [4-MCP] = 0.40mM, I = 6.45 107 einstein l1 s1, and = 250nm at pH 2.5 and 9.5 [233]. It can be observed a linear correlation, which indicates that kinetics of the degradation of these CPs by directUV-photolysis follows a pseudo-first order kinetics.

works differ in the way hydrogen peroxide is supplied tothe system. Among them, the ones that use an external sup-ply [82,197,241] and with an electrochemical production(electro-Fenton) [35,36,115] are of particular interest. On the


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Fig. 5. (Continued).

other hand, Fenton/ultrasound technologies have also beensurveyed [151].

As a result of the mineralization or degradation of CPs intointermediates, the pre-oxidized effluent is generally moreamenable to conventional treatment, e.g. flocculation andbio-treatment. Significantly, this may be achieved with ahydrogen peroxide dose of 5075% of the stoichiometry[242]. However, small chlorinated alkanes, n-paraffins andshort-chain carboxylic acids show resistance to oxidation[28]. On the other hand, removal of pollutants in turbidwaters appears to be not feasible for solutions with a highabsorbance.

As reported by several authors [227,228,143], the degra-dation and dechlorination of CPs by Fentons reagentfollows a pseudo-first order kinetics. Half-life times and

pseudo-first kinetic constants for degradation of CP andTOC are summarized in Table 5. As indicated previously byBarbeni et al. [16], the degradation rates of MCPs decreasein the order 3-MCP > 2-MCP > 4-MCP for the same ex-perimental conditions, which seems to be opposite to theones related to CP biodegradability, ortho > para > meta[62,246]. This result may suggest that OH and Cl sub-stituents at ortho and para positions on the aromatic ringinhibit the organic oxidation. Half-life times and kinetic

Fig. 6. Evolution of quantum yield () with pH for 4-MCP, 2,4-DCP, 2,4,6-TCP and 2,3,4,6-TTCP. [CP] = 0.30mM, I= 3.52 105 einstein l1 s1,and = 185436nm [21].

constants for the degradation of 2,4,5-TCP are identical tothose for 4-MCP under the same experimental conditions,respectively. On the other hand, as reported in Table 5 andshown in Fig. 7, the degradation of different CPs is higherfor series 2,3,4,6-TTCP < 2, 4, 6-TCP < 2, 4-DCP 0.4 maybe considered thoroughly biodegradable. Acetic acid andphenol are hardly biodegradable, but 4-CP and 2,4-DCP arerefractory to biological treatments.

5.1.2.3. Effect of catalytic ferrous ion concentration [Fe2+].

In the absence of iron, there is no evidence of radical orother intermediate active species formation when, for ex-ample, hydrogen peroxide is added to a wastewater whichcontains CPs. As the concentration of iron is increased, CPremoval accelerates until a point is reached where furtheraddition of iron becomes inefficient. The feature of an opti-mal dose range for iron catalyst is characteristic of Fentons


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Fig. 16. Evolution of 1/kCP with respect to 1/[H2O2] for the degradation of 4-MCP with Fentons reagent. (a) [4-MCP] = 2.00mM, [Fe2+] = 0.30mM,pH 3.0 [143]. (b) [4-MCP] = 2.33mM, [Fe2+] = 8.20 mM, pH 4.0 [41].

reagent, although the definition of the range varies betweencontaminated waters. Typical ranges are 1 part iron per 525parts of hydrogen peroxide (w/w) [242] [263].

For most applications, it does not matter whether fer-rous (Fe2+) or ferric (Fe3+) ions are used to catalyze thereactionthe catalytic cycle indicated in Fig. 14 beginsquickly if hydrogen peroxide and organic pollutants areabundant. However, if low doses of Fentons reagent are used(e.g.


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Fig. 17. Evolution of 1/kCP with respect to 1/[Fe2+

] for the degradation of 4-MCP with Fentons reagent. [4-MCP] = 2.00 mM, [Fe2+

] = 0.30mM, pH3.0 [143].

temperatures between 10 and 40 C:

k = 1.84 1010 exp

8267

T

(min1, T in K,

[TCP]0 = 0.50mM, [H2O2]0 = 5.0 mM,

[TCP]0 = 0.10mM, = 185436 nm) (19)

Fig. 18. Evolution ofkCP with respect to [CP] for the degradation of 4-MCP with Fentons reagent ([H2O2] = 6.00 mM, [Fe2+] = 0.30 mM, pH 3.0) [143].

5.2. Fenton-like reagent s (H2O2/Fe2+-solid)

Even though the degradation of CPs with Fentons reagentappears to be promising, its disadvantage is that the ho-mogeneous catalyst, added as iron salt, cannot be retainedin the degradation process. In some early attempts made,


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homogeneous catalysts are replaced by heterogeneous metalsupported ones [5,73,153,236]. The catalytic activity can berelated to hydroxyl radicals generated from hydrogen perox-ide activated by iron ions simultaneously leached from thesupport material, which act homogeneous catalysts.

Among the various catalysts studied, goethite (-FeOOH)

is one of the most efficient catalysts for Fenton-like degra-dation of CPs due to the catalysis of its surface and fer-rous ion generation. It is thought to be suitable in removalof hazardous pollutants as it exists in soil and can be recy-cled to further use, as has been reported by several authors[153,264]. Ferrous ions can be regarded to be produced fromthe reductive dissolution of goethite shown below [258]:

-FeOOH(s) + 3H++ e Fe2+ + 3H2O (20)

On the other hand, the following equation provides electrons:

H2O2 2H++ O2 + 2e

(21)

Combining both equations, the following equation is ob-tained:

2-FeOOH(s) + 4H++ H2O2 2Fe

2++ O2 + 4H2O

(22)

Hydroxyl radicals are therefore produced by Fenton reac-tion:

Fe2+ +H2O2 Fe3++ HO + HO (23)

As it is widely assumed, the dissolution of goethite playsan important role in the goethite/H2O2 Fenton-like pro-cess. It can interact chemically with H+, HO, cations,

and anions, followed by a series of dissolution reactions.

Table 6Half-life times and pseudo-first kinetic constants for degradation of CPs by Fenton-like reagent for different initial concentrations of CP, H2O2 and Fe2+

charge on the surface of the solid at pH 3.0

CP congener Solid A (m2 g1)(BET)

dp (mm) [Solid]a

(g l1)[CP]0(mM)

Total Fe(gkg1)

[H2O2]0(mM)

CP t1/2(min)

kCP (min1) Reference

2-MCPb GAC-Fe 1.0 5.5 280 426 2.27 103 [119]2-MCP Goethite 194215 0.0440.21 0.2 0.39 2.20 174 2.47 102 [161]2-MCP Goethite 194215 0.0440.21 0.2 0.39 c 2.20 36 2.75 102 [161]2-MCP Goethite 194215 0.0440.21 0.2 0.39 c 2.20 30 2.24 102 [161]2-MCP Goethite 208 0.0800.15 0.2 0.31 1.96 252 1.33 102 Lu et al. (2002)

4-MCP Iron powder 70100 1.0 7.78 155.8 1.5 1.40 10

2 [163]4-MCP Graphite-Fe


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rate of photolysis of aqueous hydrogen peroxide is about 50times slower than ozone. This technique requires a relativelyhigh dose of H2O2 and/or a much longer UV-exposure timethan, for example, UV/O3 process. On the other hand, therate of photolysis of hydrogen peroxide has been found to bepH dependent and increases when more alkaline conditions

are used, because, at 253.7 nm, peroxide anions HO2

maybe formed, which display a higher molar absorptivity thanhydrogen peroxide, 240 M1 cm1 [86], by the followingreaction:

HO2+ h HO + O (25)

The absorptivity of hydrogen peroxide may be increased byusing UV lamps with output at lower wavelengths. The mainreactions of the system are the following ones [22,88]:

Initial reaction:

H2O2 + hkp

kp

2HO

H2O2 H2O+ 12 O2

Propagation:

HO + H2O2 HO2 + H2O

HO2 +H2O2 HO + H2O+ O2

HO2 +HO2 HO + HO + O2

Termination:

HO + HO2 H2O+ O2

HO + HO H2O2 + O2

CP degradation:

CP+ HOkrIntermediates,

CP+ H2O2 Intermediates+ H2O,

CP+ h Intermediates (26)

Intermediates+H2O2 CO2 + H2O+ Cl

,Intermediates+HO CO2 + H2O+ Cl

,

Intermediates+ h CO2 + H2O+ Cl (27)

If there are other strong UV absorbers different than hy-drogen peroxide, the observed effect will be the same as ifthe incident flux were decreased, i.e. there is less radiationintensity available for the photolysis of hydrogen peroxide,or the amount of radiation transformed into HO radicals willbe lower if such absorbents are present. Furthermore, CPscan be hydroxylated by the photolysis of hydrogen peroxide[158,217,218]. However, Hivronen et al. [108] reported that

rates achieved by the H2O2/UV process in alkaline condi-tions with DCP were comparable to those achieved by di-rect DCP irradiation, while at acidic pH the addition of hy-drogen peroxide increased the reaction rate by one order ofmagnitude.

Ultimate oxidation of CPs to carbon dioxide and water

has rarely been obtained under typical test conditions. Assummarized in Table 7, typical values for half-life times arebetween 0.3 and 20.1 min for CP degradation, depending onthe initial concentration of CP and hydrogen peroxide, theintensity of radiation and the degree of chlorination. It is ob-served that the degradation rates increase when the numberof chlorine substituents decreases. On the other hand, dataconcerning with pseudo-first order kinetic constants are alsoincluded. As reported by Bentez et al. [23], this rate con-stants present moderate higher values than those obtained inthe single photolytic process (see Table 5), which demon-strate an additional contribution to the photolytic reaction,which can be related to the action of hydroxyl radicals gen-

erated by the presence of hydrogen peroxide.According to Bentez et al. [22,23], the total kinetic con-

stant, kt, may be assumed to be constituted by two contri-butions (kt = kp+ kr). Some constants related to the secondreaction (kr) are summarized in Table 8. It can be observedthat these constants tend to decrease when the degree ofchlorination increases. Specifically, Sundstrom et al. [228]proposed for the ratio kt/kp values of 2.0 and 1.3 for 2,4-DCPand 2,4,6-TCP, respectively.

5.4. Photo-Fentons reagent (H2O2/Fe2+/UV)

The photo-Fenton reaction is also well-known in theliterature [134,207,209], which is an efficient and in-expensive method for wastewater and soil treatment[18,48,132,138,206,210]. Photo-Fenton is known tobe able to improve the efficiency of dark Fenton orFenton-like reagents, respectively, by means of the in-teraction of radiation (UV or Vis) with Fentons reagent[107,196,226,221,265]. This technique has been suggestedto be feasible and promising to remove pollutants from nat-ural and industrial waters and increase the biodegradabilityof CPs being used as a pre-treatment method to decrease thetoxicity of water [74,75,167,181]. Some innovative applica-tions dealing with photo-Fentons reagent include oxalateas a ligand of iron ions [8].

As summarized in Table 9, typical values for half-lifetimes are between 0.3 and 50 min for CP degradation, de-pending on the initial concentration of CP, hydrogen perox-ide and ferrous ions, the intensity of radiation and the degreeof chlorination. It is observed that the degradation rates in-crease when the number of chlorine substituents decreases.As reported by Bentez et al. [23], pseudo-first kinetic con-stants this rate present moderate higher values than thoseobtained in the single photolytic process (see Table 5), asindicated for UV/H2O2 reagent. On this way, an additionalcontribution to the photolytic reaction may be considered,


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Table 7Half-life times and pseudo-first kinetic constants for degradation of CPs by H 2O2/UV reagent for different initial concentrations of CP and H2O2 andfor different radiation intensities

CP congener [CP]0(mM)

[H2O2]0(mM)

pH T (C) Rad. intensity(einstein l1 s1)

Wavelengthrange (nm)

CP t1/2 (min) kCP (min1) Reference

2-MCP 0.40 40.00 2.5 Room 6.45E07 250 5.07 102 [233]2-MCP 0.40 40.00 9.5 Room 6.45E07 250 5.11 102 [233]

4-MCP 0.30 0.50 2.0 25 3.52E05 185436 0.3 0.549 [21]4-MCP 10.00 30.00 3.5 20 1.04E06 320400 Not reached [257]4-MCP 0.30 0.50 2.0 Room 4.80E04 185436 1.0 0.601 [23]4-MCP 0.40 40.00 2.5 Room 6.45E07 250 9.9 5.07 102 [233]4-MCP 0.40 40.00 9.5 Room 6.45E07 250 10.1 5.11 102 [233]4-MCPa 1.00 10.00 3.0 25 3.34E04 1420 2.80 102 [17]

2,4-DCP 0.30 0.50 2.0 Room 4.80E04 185436 14.5 4.40 102 [23]2,4-DCP 0.30 0.50 2.0 25 4.80E04 185436 2.54 102 [21]2,4-DCP 0.40 40.00 2.5 Room 6.45E07 250 7.9 8.64 102 [233]2,4-DCP 0.40 40.00 9.5 Room 6.45E07 250 11.0 5.22 102 [233]2,4-DCP 0.74 0.74 25 4.87E04 240570 110.0 [118]

2,4,6-TCP 0.30 0.50 2.0 Room 4.80E04 185436 19.2 3.30 102 [23]2,4,6-TCP 0.30 0.50 2.0 25 4.80E04 185436 1.80 102 [21]2,4,6-TCP 0.40 40.00 2.5 Room 6.45E07 250 7.7 0.109 [233]

2,4,6-TCP 0.40 40.00 9.5 Room 6.45E07 250 16.5 5.46E02 [233]

2,3,4,6-TTCP 0.30 0.50 2.0 Room 4.80E04 185436 20.1 2.90 102 [23]2,3,4,6-TTCP 0.30 0.50 2.0 25 4.80E04 185436 2.90 102 [21]

PCP 0.04 40.00 2.5 Room 6.45E07 250 0.458 [233]PCP 0.04 40.00 9.5 Room 6.45E07 250 0.267 [233]

a Data referred to TOC.

which can be related to the action of hydroxyl radicals gen-erated by the presence of hydrogen peroxide.

As pointed out by Bentez et al. [22,23], in the degrada-tion of CPs by photo-Fentons reagent, additionally to thephotolysis of hydrogen peroxide, the production of hydroxylradicals by the Fentons reagent must be taken into account.On this way, it must be expected that the hydroxyl attack re-action must improve the removal obtained by the UV/H2O2reagent, because in this case two different pathways con-tribute to the generation of free radicals and, therefore, theirconcentration must be higher. All initial, propagation, ter-mination and CP degradation steps for CP removal are thesame as the ones exposed for reactions with Fentons andUV/H2O2 reagents. Nevertheless, another step for initial re-actions must be added [48,242,243], which does not takeplace in Fenton reaction, because there is no radiation:

Initial reactions:

H2O2 + Fe2++ h Fe3+ + HO + HO (28)

Table 8Values for kt and kr pseudo-first kinetic constant for the degradation ofCP with H2O2/UV reagent [22,23]

CP kt 103 kr 103

4-MCP 601 36.02,4-DCP 44 6.92,4,6-TCP 33 7.22,3,4,6-TTCP 29 7.9

According to Benitez et al. [22,23], as has been pointed forthe H2O2/UV process, the total kinetic constant, kt, maybe assumed to be constituted by two contributions (kt =kp+kr). Some constants related to this reaction (kr) are sum-marized in Table 10. It can be observed that these constantstend to decrease when the degree of chlorination increases.Specifically, Sundstrom et al. [226] proposed for the ratiokt/kp values of 2.0 and 1.3 for 2,4-DCP and 2,4,6-TCP, re-spectively.

Due to this additional generation of hydroxyl radical in thedegradation of CPs by the photo-Fenton reaction, the degreeof degradation is higher compared to direct photolysis ordegradation with UV/H2O2 reagent for the same system ofreaction, as shown in Fig. 19. In fact, this improvement ismore important for chlorinated CPs instead of MCPs, whichshow a similar degradation for the three oxidation processes.

6. AOPs based on ozonation

6.1. Ozone (O3)

There has been an increasing interest in the last decadesin using ozone to treat effluents containing hazardous pol-lutants with the development of large-scale ozone gener-ators along with reduced installation and operating costs[14,96,176,189,199,201,202]. Compared to other oxidiz-ing reagents, ozonated water is more efficient in pollutantdegradation and it is not harmful for most of the organisms,


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Table 9Half-life times and pseudo-first kinetic constants for degradation of CPs by photo-Fentons reagent for different initial concentrations of CP and H 2O2and for different radiation intensities

CP congener [MCP]0(mM)

[H2O2]0(mM)

[Fe2+]0(mM)

pH T (C) Rad. intensity(einstein l1 s1)

(nm) CP t1/2 (min) kCP (min1) Reference

4-MCP 0.30 0.50 0.010 2.0 25 3.52E05 185436 0.25 0.642 [21]4-MCP 10.00 30.00 0.005 3.5 20 1.04E06 320400 22.0 [257]

4-MCP 0.80 4.00 0.020 2.0 25 7.30E02 >300 28.0 [220]4-MCP 0.30 0.50 0.010 2.0 Room 3.76E04 185436 0.9 0.642 [23]4-MCP 1.00a 10.00 0.25 3.0 25 3.34E04 13.0 1.25 [17]

2,4-DCP 0.80 4.00 0.020 2.0 25 No data >300 25.0 [220]2,4-DCP 0.30 0.50 0.010 2.0 Room 4.80E04 185436 0.9 8.80 102 [23]2,4-DCP 0.30 0.50 0.010 2.0 25 3.52E05 185436 8.80 102 [21]

2,4,5-TCP 0.80 4.00 0.020 2.0 25 No data >300 Not reached [220]2,4,6-TCP 0.30 0.50 0.010 2.0 Room 4.80E04 185436 8.5 7.80 102 [23]2,4,6-TCP 0.30 0.50 0.010 2.0 25 3.52E05 185436 7.80 102 [21]

2,3,4,6-TTCP 0.30 0.50 0.010 12.0 Room 4.80E04 185436 12.0 5.80 102 [23]2,3,4,6-TTCP 0.30 0.50 0.010 25 3.52E05 185436 5.80 102 [21]

PCP 0.80 4.00 0.020 50 25 No data >300 50.0 [220]

a

This CP concentration corresponds to TOC = 73mgl1

.

Table 10Values for k and kr pseudo-first kinetic constant for the degradation ofCP with photo-Fentons reagent

CP kt 103 kr 103

4-MCP 642 16.02,4-DCP 88 56.82,4,6-TCP 78 66.72,3,4,6-TTCP 58 63.8

because no strange compounds are added to treated wa-ters. Ozonation has been widely used for drinking waterdisinfectionbacterial sterilization, odor, algae, and tri-halomethane removal and organic compound degradation[2,91], but its application to wastewater treatment is lim-ited due to its high energy demand. Thanks to its oxidizingpower and the absence of hazardous decomposition prod-ucts, ozone is a potential pre-treatment agent to transformrefractory compounds into substances that can be further re-moved by conventional methods [13,116]. Thus, the ozona-

Fig. 19. Degradation of CPs by different AOPs ([CP] = 0.30mM, [H2O2] = 0.50 mM, [Fe2+] = 0.010mM, pH 2.0) [21,23].

tion of dissolved compounds in water can constitute anAOP by itself, as hydroxyl radicals can be generated fromthe decomposition of ozone, which might be catalyzed byhydroxyl ions or initiated by the presence of traces of othersubstances, like transition metal cations [224]. As pH in-creases, so does the rate of decomposition of ozone in water.

The reaction of aqueous phenol with ozone hasbeen extensively studied and the kinetics and manyof the products of the reaction have been determined[2,9,111,112,150]. The specific application of ozonation

to degrade CPs has been also studied by several authors[31,49,63,81,121,127,142,233]. As reported by several au-thors [23,143,188,233], the degradation and dechlorinationof CPs by ozone reagent follows a pseudo-first order ki-netics. Half-life times and pseudo-first rate constants fordegradation of CP and TOC are summarized in Table 11.

The degradation of CPs is favored at high pH [22,101,140,141,233]. The faster degradation of this compounds in alkalimedia can be due to the fast production of HO radicalsand the dissociation of CPs to chlorophenolate ions that are


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Table 11Half-life times and pseudo-first kinetic constants for degradation of CPs by ozone reagent for different initial concentrations of CP

Pollutant [CP]0 (mM) O3 feed flowrate (mgmin1)

pH CP t1/2(min)

kCP (min1) kO3 (min1) Reference

2-MCPa 0.78 5.0 6.0 17.0 1.49 103 [140]2-MCP 0.40 107.3 2.5 14.0 4.78 102 [233]2-MCP 0.78 18.0 3.0 5.60 102 [188]

2-MCP 0.78 18.0 7.0 7.5 8.69 102b [188]2-MCP 0.78 18.0 9.0 0.137 [188]2-MCP 0.18 8.3 3.4 28.8 [1]4-MCP 0.30 2.0 38.5 1.70 102 [23]4-MCP 0.30 9.0 3.4 0.239 [23]4-MCP 0.10 2.0 8.3c 4.48 [22]4-MCP 0.30 2.0 3.6c 11.3 [22]4-MCP 0.40 2.0 2.4c 13.9 [22]4-MCPa 0.78 6.0 6.0 2.30 103 [140]4-MCP 0.40 107.3 2.5 19.0a 3.64 102 [233]4-MCPa,d 1.38 8.3 3.3 7.0, 22.0a 0.163, 1.19 102 a [210]4-MCP 0.18 8.3 3.4 16.2 [1]4-MCPe 0.50 ? 7.0 12.7, 170.0a 2.55 [98]

2,4-DCP 0.30 2.0 30.4 2.40 102 [23]

2,4-DCP 0.30 9.0 3.3 0.315 [23]2,4-DCP 0.40 2.0 45.3 [22]2,4-DCPa 0.78 5.0 6.0 6.0 3.67 103 [140]2,4-DCP 0.18 8.3 3.4 22.4 [1]2,4-DCP 1.2 7.0 300.0 1.07 103 [266]

2,4,6-TCP 0.30 2.0 20.6 4.40 102 [23]2,4,6-TCP 0.30 9.0 3.1 0.314 [23]2,4,6-TCP 0.40 2.0 228.0 [22]2,4,6-TCP 0.18 8.3 3.4 156.0 [1]2,4,6-TCP 0.16 ? 4.9 0.552 [55]

2,3,4,6-TTCP 0.30 2.0 10.6 9.40 102 [23]2,3,4,6-TTCP 0.30 9.0 1.9 0.415 [23]2,3,4,6-TTCP 0.40 2.0 361.0 [22]

a Data referred to TOC.b Half-time for TOC not reached.c Half-times in seconds.d Data referred to TOC = 100mgl1.e Data referred to TOC = 36mgl1.

able to react with ozone faster than non-dissociated species[111]. However, according to the results of Hautaniemi et al.[101], reactions with HO radicals did not appear to makeany contribution to the rate of oxidation at pH 9, which wasrelated to the high rate of the reaction between phenolate ionswith ozone at pH conditions which usually favor oxidationby means HO radicals.

Not much data are present in the literature for the degra-dation of CPs. Otherwise, an interesting work concerningthis subject was carried out by Bentez et al. [22], who stud-ied the ozonation process of several CPs in alkali media andfound that the degradation curves were very close for allCPs. Pseudo-first order kinetic constants and half-life timesalso confirmed the small differences in the removal of theCPs at this pH as depicted in Fig. 20. A similar conclusionwas reached by Hong and Zeng [113], who found that therates of disappearance of PCP were very similar at pH be-tween 7 and 12, reflecting a small influence of pH.

6.1.1. Mechanistic pathway for ozonation

One of the most important facts to mention when study-ing the oxidation of CP by ozone is the high influence ofthe pH in the kinetics and pathways of the reaction. Thisarises from the fact that pH affects the double action ofozone on the organic matter, that may be a direct or an in-direct (free radical) ozonation pathway [20,110], as shownin Fig. 21. These different reaction pathways lead to dif-ferent oxidation products and are controlled by differentkinetic models. At low pH, ozone exclusively reacts withcompounds with specific functional groups through selec-tive reactions such as electrophilic, nucleophilic or dipolaraddition reactions (i.e. direct pathway) [144]. On the otherhand, at basic conditions, ozone decomposes yielding hy-droxyl radicals, which are high oxidizing species [147] thatreact in a non-selectively way with a wide range of organicand inorganic compounds in water (i.e. indirect ozona-tion) [37]. Normally, under acidic conditions (pH


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Fig. 20. Decomposition curves for the ozonation of 4-CP, 2,4,6-TCP and 4-CGC (4-chloroguaiacol; 4-chloro-2-methoxy phenol) in experiments at pH 9[22].

direct ozonation dominates, in the range of pH 49 both arepresent, and above pH >9 the indirect pathway prevails.

6.1.1.1. Direct ozonation. When the ozonation is carriedout in acidic media (pH 2) the decomposition of ozone,which is initiated by the action of HO ions [225], is muchtoo low and consequently the formation of hydroxyl radi-cals is limited. According to this, the main pathway in thedegradation of CPs at these conditions is the direct ozona-tion, which is a selective reaction with quite slow kineticconstants, which lie in the range of 11000M1 s1.

Trapido et al. [233] have reported that an increase in the

number of chlorine atoms in the aromatic ring provides an in-crease in the degradation rate, since the presence of chlorineenhances the dechlorination step, and therefore, the degra-dation is faster. This statement agrees with the results ob-tained by Bentez et al. [22]. The results obtained, shown inFigs. 22 and 23, indicate that the higher is the pH, the fasterare the degradation of these CPs. In addition, the slowestdegradation when the ozonation was undertaken at pH 2 and4 belonged to 4-CP, whereas no significant differences be-tween the three compounds were found at the highest pH.Similar results were found by Kuo et al. (1999), who stud-ied the ozonation of 2-CP, 4-CP and 2,4-DCP at pH 6 and 9.After 10 min of reaction at pH 6 the degradation of 2,4-DCP,

4-CP and 2-CP were 79.2, 63.9 and 34.3%, respectively.

6.1.1.2. Indirect ozonation. The indirect reaction pathwayinvolves radicals. The first step is the decay of ozone, accel-erated by initiators HO ions, to form secondary oxidantssuch as HO radicals, which react non-selectively and im-mediately (k = 1081011 M1 s1) with solutes. The radi-cal pathway is very complex and influenced by many sub-stances. The mechanism can be divided in three differentparts:

Initial reactions:

O3 +HO O2+HO2 , k1 = 70 M

1 s1 (29)

Propagation:

O3 + O2 O3

+O2, k2 = 1.6 10

9 M1 s1

(30)

O3+ H+ HO3, pKa = 6.2 (31)

HO3 HO +O2, k3 = 1.1 108 M1 s1 (32)

HO +O3 HO4, k4 = 2.0 109 M1 s1 (33)

HO4 HO2 + O2, k5 = 2.8 104 s1 (34)

HO2 O2+ H+, pKa = 4.8 (35)

Termination:

HO + CO32 HO + CO3

,

k6 = 4.2 108 M1 s1 (36)

HO +HCO3 HO + CO3,

k7 = 1.5 107 M1 s1 (37)

CP degradation:

CP +HO Intermediates,

CP +O2 CPOO Intermediates+ HO2 (38)

Intermediates+ HO CO2 +H2O+ Cl,

Intermediates+ O2 IOO CO2 + H2O+ Cl (39)

6.2. O3/UV reagent

Photolytic ozonation (O3/UV process) is an effectivemethod for the oxidation and destruction of toxic and refrac-tory organics in water [20,55,72,86,130,194,198,206,267269] and has a significant potential as a wastewater treat-


24/38


Fig. 21. General scheme of reactions for ozonation of CPs in an aqueous solution, including the effect of radiation and H2O2, in case they are employed.The final products of degradation should be CO2, H2O and Cl if complete mineralization is achieved [111].

Fig. 22. Decomposition curves for the ozonation of 4-CP, 2,4,6-TCP and 4-CGP in experiments at pH 2 [22].


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Fig. 23. Decomposition curves for the ozonation of 4-CP, 2,4,6-TCP and 4-CGP in experiments at pH 6 [22].

ment process. Basically, aqueous systems saturated withozone are irradiated with UV light of 253.7 nm. The extinc-

tion coefficient of ozone at 253.7 nm is 3300 M1

cm1

,much higher than that of hydrogen peroxide. The decay rateof ozone is around 1000 times higher than that of hydrogenperoxide [94]. It was shown that photolytic ozonation ismore effective for the destruction of some organic com-pounds than either UV-photolysis or ozonation alone. How-ever, some researchers [148] have found photolytic ozona-

Table 12Half-life times and pseudo-first kinetic constants for degradation of CPs by O 3/UV reagent for different initial concentrations of CP and O3 flow ratesand for different radiation intensities

Pollutant [MCP]0(mM)

O3 feed flowrate (mgmin1)

pH Rad. intensity(einstein l1 s1)

(nm) CP t1/2 (min) kCP (min1) References

2-MCPa 0.78 5.0 6.0 5.97 105 320400 15.0 7.68 103 [140]2-MCP 0.40 107.3 2.5 6.45 107 250 3.83 102 [233]2-MCP 0.18 8.3 3.4 2.50 105 9.9 [1]4-MCP 0.30 2.0 1.76 105 185436 1.3 [23]4-MCPa 0.78 6.0 5.97 105 320400 4.5 1.18 102 [140]4-MCP 0.40 107.3 2.5 6.45 107 250 3.65 102 [233]4-MCPb 1.38 8.3 3.3 1.38 108 7.2, 46.0a 0.112, 5.42 102 a [210]4-MCP 0.18 8.3 3.4 2.50 105 9.3 [1]4-MCPa,c 0.78 7.0 4.33 105 320400 10.0 5.20 103 [141]4-MCP 0.18 7.0 2.56 104 172 10.5, 92a 3.87 [98]

2,4-DCP 0.30 2.0 1.76 105 185436 15.6 6.50 102 [23]2,4-DCPa 0.78 5 6.0 5.97 105 320400 8.8 8.59 103 [140]2,4-DCP 0.40 107.3 2.5 6.45 107 250 5.16 102 [233]2,4-DCP 0.18 8.3 3.4 2.50 105 16.8 [1]

2,3,5-TCP 0.04 3.0 8.08 105 [239]2,3,6-TCP 0.12 3.0 8.08 105 [239]2,4,6-TCP 0.30 2.0 1.76 105 185436 11.8 6.80 102 [23]2,4,6-TCP 0.18 8.3 3.4 2.50 105 39.5 [1]2,4,6-TCP 0.24 3.0 8.08 105 [239]2,4,5-TCP 6.0 3.0 8.08 105 [239]2,4,5-TCP 0.18 8.3 3.4 2.50 105 25.0 [1]

2,3,4,6-TTCP 0.30 2.0 1.76 105 185436 9.7 7.10 102 [23]2,3,4,6-TTCP 0.05 3.0 8.08 105 [239]

PCP 0.05 3.0 8.08 105 [239]

a Data referred to TOC.b Data referred to TOC = 100mgl1.c Data referred to TOC = 54.6 m g l1.

tion to be only more effective than ozonation alone in somecases.

The photodegradation of CPs by ozone and UV radi-ation, alone or combined with hydrogen peroxide or fer-rous and ferric ions, has been reported by several authors[1,27,168,206,217,218,226]. The initial concentrations ofCP, O3 flow rates, half-life times and pseudo-first kineticconstants for degradation of CPs by photo-ozonation areshow in Table 12.


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Several authors have compared the O3/UV systemwith the direct photolysis or simple ozonation. Some ofthem [23,72,139141,168] concluded that the combinationO3/UV improves the efficiency of the single processes.Kuo [140,141] compared the destruction of 2-CP, 4-CP and2,4-DCP by combination of photolysis+ ozonation, direct

photolysis and ozonation alone. As a result of this researchit was found that the synergistic effects of the combinedprocedure increased obviously with increasing initial pH ofsolution to basic levels. This fact was attributed to that theinitiation of ozonation could be catalyzed by OH ions.The experimental results also proved that the potentiali-ties of photolytic ozonation compared to direct photolysisand ozonation alone was remarkable for the treatment ofindustrial wastewaters containing CPs.

Bentez et al. [23] studied the degradation of four CPs bythe combined process UV/O3 at 25 C and pH 2. When com-paring their results with single photodecomposition or pH2 ozonation, it was found that the combined system accel-

erated the decomposition with an extremely high rate con-stant for 4-CP. The other CPs studied (2,4-DCP, 2,4,6-TCPand 2,3,4,6-TTCP) presented slightly increasing rate con-stants when the substituent chlorine atoms increased (thissequence was also observed in the single ozonation). Esplu-gas et al. [72] applied the photolytic ozonation to degrade4-CP. It was concluded that UV on its own produced neg-ligible degradation of the compound and some degradationwas achieved by single ozonation. The mineralization of4-CP was more effective when it was subjected to photolyticozonation.

Despite the conclusions reached by the above mentioned

authors, Trapido et al. [233] affirmed that although the mo-lar absorptivity of CPs in known to be relatively high in theUV, the combination of UV with ozone did not acceleratethe degradation of CPs further. In this research, the degrada-tion of 4-CP, 2-CP, 2,4-DCP, 2,4,6-TCP, 2,3,4,6-TTCP andPCP was studied by several AOPs. From this study, it wasconcluded that CPs can be degraded at quite significant ratesby all the methods studied, while traditional ozonation athigh pH was determined to be the most effective method totreat CPs.

6.2.1. Mechanistic pathway

The degradation of CPs by O3/UV reagent follows amechanism similar to the one described for simple ozona-tion. Nevertheless, a synergistic effect between directozonation, direct photolysis and hydroxyl radical decom-position is observed in this system [23]. On this way, someother elemental stages have to be added to take into ac-count the interaction between radiation and matter (seeFig. 21). First of all, a stage the photolysis of ozone mustbe added, which leads to two hydroxyl radicals, which donot act as they recombine producing hydrogen peroxide[194]:

H2O+O3 + h H2O2 + O2 (40)

Hydrogen peroxide, on being exposed to UV radiation canthen decompose directly into hydroxyl radicals:

H2O2 + h 2HO (41)

In addition, H2O2 accelerates O3 decomposition into hy-droxyl radicals [225]:

2O3 +H2O2 2HO + 3O2 (42)

6.3. Ozonation+ homogeneous/heterogeneous catalysis:

O3/H2O2, O3/Fe2+ and O3/Fe

2+/UV

Some approaches have been taken into account in order toimprove the oxidizing power of ozone or ozone/UV leadingto the reduction of time required for the reaction and hence,decreasing its energy cost. Several works indicate that acombination of ozone with hydrogen peroxide (O3/H2O2),which actually acts as a homogeneous catalyst, improves CPdegradation [3,192,210,270]. In this case, in addition to thegeneral pathway for ozonation exposed in Fig. 21, ozonereacts with H2O2 when present as anion HO2 and the directreaction of ozone with the undissociated H2O2 is negligibledue to a very low kinetic constant:

H2O2 + O3 H2O+ O2, k < 0.01 M1 s1 (43)

Other propagation reactions that must be added are the fol-lowing ones:

H2O2 HO2 + H+ (44)

O3 +HO HO2

+ O2 (45)

H2O2 + H2O HO2+H3O

+ (46)

O3 +HO2 HO + O2

+ O2 (47)

O3 +HO HO2 +O2 (48)

In absence of any contaminant, the chain reaction may beterminated by the major step below:

HO + HO2 H2O+O2, k = 3.7 1010 M1 s1

(49)

In most cases, the optimum ratio H2O2/O3 is considered tobe 0.5 mol/mol or more. Chamarro et al. [41] found the sto-ichiometric coefficients for 4-MCP degradation in the range0.6010.044 (mol removed/mol active reagent). Despite thegood results shown by this method, it is not always advanta-geous to add H2O2 with O3, even when HO radical chem-istry is necessary for the oxidation of the target organic. Onthe other hand, promising experiments involving ferrous orferric ion homogeneous catalysis together with ozonation inpresence or in absence of radiation are also reported in theliterature [2,271]. Some data concerning this processes areshown in Table 13. As indicated in Fig. 24, the presenceof ferrous ions seems to improve the mineralization of CPs


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Table 13Half-life times and pseudo-first kinetic constants for degradation of CPs by O 3/Fe2+/UV reagent for different initial concentrations of CP and O3 flowrates and for different radiation intensities

Pollutant [CP]0 (mM) [Fe2+]0(mM)

O3 feed flow rate(mgmin1)

pH Rad. intensity(einstein l1 s1)

(nm) CP t1/2 (min) kCP (min1) References

2-MCP 0.18 0.067 8.3 3.4 2.50 105 7.1 [1]4-MCP 1.38 1.00 8.3 3.3 1.38 108 7.0, 22.0a 0.113, 6.42

102 a

[210]

4-MCP 0.18, 100a 0.067 8.3 3.4 2.50 105 5.8 [1]

2,4-DCP 0.18 0.067 8.3 3.4 2.50 105 12.4 [1]

2,4,6-TCP 0.18 0.067 8.3 3.4 2.50 105 19.0 [1]2,4,5-TCP 0.18 0.067 8.3 3.4 2.50 105 10.5 [1]

a Data referred to TOC. Concentrations are measured in mg l1.

Fig. 24. Evolution of Ln[TOC] with time for the degradation of 4-CP in presence of O3/Fe2+/UV reagent ([4-CP] = 0.18mM, [TOC]0 = 100mgl1,[Fe2+]0 = 1.00 mM, pH 3.3, I= 1.38 108 einstein l1 s1) [210].

Table 14Half-life times and pseudo-first kinetic constants for degradation of CPs by heterogeneous ozonation [188]

Solid a (m2 g1) (BET) dp (mm) Conc. susp. (g l1) [2-CP]0 (mM) pH CP t1/2 (min) TOC t1/2 (min) kCP (min1)

-Alumina 120190 0.0600.200 1.7 0.78 3.0 0.0653-Alumina 120190 0.0600.200 1.7 0.78 7.0 7.0 79 0.1269-Alumina 120190 0.0600.200 1.7 0.78 9.0 0.1390

when it is compared to single ozonation, and in presence ofUV radiations, the action of these ions also seems to im-prove the efficiency O3/UV process.

On the other hand, Ni and Chen [188] have reported theuse of heterogeneous catalytic ozonation of 2-CP by employ-ing -alumina as a solid catalyst, as summarized in Table 14.The research showed that the rate of degradation of TOCcould increase from 21 to 43% by using this method insteadof single ozonation, and the pseudo-first reaction constantcould also increase from 0.8688 to 0.127 min1 at pH 7.0for particles with the general features exposed in Table 14.

7. Photocatalysis

Photocatalytic degradation has proven to be a promis-ing technology for degrading refractory chlorinated aromat-

ics [38,50,56,212,232] and more than 1700 references havebeen recently collected on this discipline [30]. Comparedwith other conventional chemical oxidation methods, pho-tacatalysis may be more effective because semiconductorsare inexpensive and capable of mineralizing various refrac-tory compounds [139], however this technique is still in thedevelopmental stage for many cases [59].

7.1. Mechanistic pathways

Photocatalytic reactions occur when charge separationare induced in a large bandgap semiconductor by exci-tation with ultra bandgap radiations [203]. This way, theabsorption of light by the photocatalyst greater than itsbandgap energy excites an electron from the valence bandof the irradiated particle to its conduction band, produc-ing a positively charged hole in the valence band and


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Fig. 25. Conduction and valence bands and electronhole pair generation in semiconductors.

an electron in the conduction band [122] as shown inFig. 25.

The hole in the valence band may react with water ab-sorbed at the surface to form hydroxyl radicals and on theother hand, the conduction band electron can reduce ab-

sorbed oxygen to form peroxide radicals anions that can fur-ther disproportionate to form HO through various pathways[4]. During the photocatalytic process, other oxygen con-taining radicals are also formed including superoxide radicalanion and the hydroperoxide radical [59]. In addition, theband electron may also react directly with the contaminantvia reductive processes [125]. It is not clear under whichexperimental conditions, one reaction pathway is more im-portant than the other. It is generally accepted that the sub-strate adsorption in the surface of semiconductors plays andimportant role in photocatalytic oxidation [272]. Thereforeif the surface of the catalyst is covered with the CP, the di-rect oxidation by positive holes could be the major oxida-tion pathway since the adsorption of organic compounds onthe surface of the catalyst is the prerequisited step for directcharge transfer. On the other hand, indirect oxidation by hy-droxyl radicals requires the adsorption of water or hydroxideions on the surface to form the hydroxyl radicals [228,229].

7.1.1. Direct photocatalytic pathway

The literature suggests two different direct photocatalyticreaction mechanisms.

7.1.1.1. The LangmuirHinshelwood process. This mech-anism is based on production of electrons and holes by the

photoexcitation of the catalyst. The hole is then trappedby the adsorbed molecule (CP) to form a reactive radicalstate. On one side, this reactive specie (CPads+) can de-cay when recombination with and electron occurs and onthe other hand its chemical reaction yields to the productsand regenerates the original state of the catalyst surface (S).Reactions (50)(55) show the reaction mechanism of theLangmuirHinshelwood process (adapted from Serpone andEmelie [215]):

M+ S CPads

(adsorption/desorption Langmuir equilibrium) (50)

CPads M+ S

(adsorption/desorption Langmuir equilibrium) (51)

Cat+ h e + h+ (photoexcitation of the catalyst)

(52)

CPads + h+ CPads

+

(hole trapping by adsorbed molecule) (53)

CPads++ e CPads (decay of the reactive state) (54)

CPads+ product+ S (chemical reaction) (55)

7.1.1.2. The EleyRideal process. This process starts withthe photogeneration of free carriers and the subsequent trap-

ping of the holes by surface defects (i.e. potential surfaceactive centers) S to produce surface active centers S+. Thesesurface active centers can then react with the CP (chemisorp-tion) to form species (SCP)+ that further decompose yield-ing the photoreaction products or can recombine with elec-trons, which represents their physical decay. The reactionsinvolved in the process are shown below (adapted from Ser-pone and Emelie [215]):

Cat+ h e + h+ (photogeneration of free carriers)

(56)

S+ h+ S+ (hole trapping by surface defects) (57)

S+ + e S (physical decay of active centers) (58)

S+ + CP (SCP)+ (chemisorption) (59)

(SCP)+ S+ products

(reaction to form photoreaction products) (60)

7.1.2. Indirect photocatalytic pathway

The radical photodegradation mechanism is shown in re-actions (61)(67) (adapted from Yue et al. [255]). The pro-cess starts with the photogeneration of electronhole pairs


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on the surface of the catalyst particles when irradiated bythe light in the region of absorption charge-transfer bands.The hole is trapped by the water molecules leading the for-mation of HO radicals and H+ and the electrons allow theformation of H2O2 which further decomposes in more OH

radicals by means of its reaction with the oxygen supplied in

the medium. Finally, the radicals formed during this mech-anism are responsible for the oxidation of the CPs yieldingsome intermediate compounds and mineralization products:

h e + h+ (61)

h+ + H2O(ads) HO(ads) + H(ads)+ (62)

O2 + 2e O2 (ads)

(63)

O2 (ads)+H+ HO2 (ads) (64)

HO2 (ads) H2O2 (ads) + O2 (65)

H2O2(ads) 2HO(ads)

HO

(66)

2HO(ads)

HO+ CPs Intermediates+HCl

CO2 + H2O (67)

7.2. Studies of degradation of CPs

Numerous photocatalytic semiconductors have beenrecently investigated for the destruction of several CPsin water, including TiO2, CdS, ZnS and active carbon[139,164,279,280]. However, titanium dioxide, has beenstudied more extensively because of its low cost, stable na-

ture and high photoactivity when exposed to near-UV light[84,109,125,139,237,203,273]. The experimental resultsobtained by Jardim et al. [124] and shown in Fig. 26 re-inforce the viability of heterogeneous photocatalysis usingTiO2 in the detoxification process, especially when chori-natedphenols are the target compounds since degradation

Fig. 26. Photodegradation of PCP, 2,3,5-TCP, 3,5-DCP and 2,4-DCP and TOC values during irradiation. Experiments carried out in a cylindricalphotoreactor of 800ml capacity with a high-pressure 125W mercury lamp, 0.1 g l1 of TiO2 and air supplied at 560 ml min1 [124].

of DCP, TCP and PCP was shown to be feasible and almostcomplete.

The applicability of some photocatalysts for CP degrada-tion is summarized in Table 15. For instance, Doong et al.[61] reported that cadmium sulfide had a higher rate con-stant than titanium dioxide for the degradation of 2-CP. This

fact was attributed to that the average particle size of cad-mium sulfide is larger than that of titanium dioxide. Otherresearchers [15] studied the photocatalytic degradation of2,4,6-trichlorophenol, 2,3-dichlorophenol, 2-chlorophenoland 2,4-dichlorophenol on aqueous suspensions of-Fe2O3and -FeOOH and TiO2. -FeOOH was found to be inac-tive for CPs degradation with exception of 2,4-DCP where amodest effect was observed. The most active photocatalysis(-Fe2O3) was compared with TiO2. Total mineralizationof CPs was observed on TiO2 while only partial mineraliza-tion was observed on Fe2O3. The effectiveness of TiO2 hasbeen also compared with Na4W10O32 [230]. The results ofthe solar degradation of 2-CP and 2,4-DCP in the presence

of TiO2 and Na4W10O32 showed that the degradation ofthe pollutants was twice faster in the presence of TiO2.

One possible way to enhance the degradation efficiencyof photocatalytic reactions consist of adding heavy metals,such as platinum (Pt), palladium (Pd), gold (Au) or othersemiconductors [47,182,183]. The increase in the photoac-tivity when metallic ions are present can be justified by tak-ing into account that the presence of a photoreduced metalon TiO2 would improve charge separation. Rideh et al. [203]have reported that the 2-CP photodegradation rate increasesremarkably when they are saturated by oxygen and metallicions such as Ag+, Cr2O72, Cu2+ and Fe2+. Chen et al. [46]

studied the photocatalytic degradation of 2-CP, 2,4-DCP and2,4,6-TCP in the presence of manganese ions at pH 3. Theresults indicate that dissolved manganese ions can increasethe rate of CPs oxidation in the TiO2 photocatalytic system.

Another technique to improve the degradation of CPsby photocatalysis consists of coupled semiconductors


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Table 15Half-life times and pseudo-first kinetic constants for degradation of CPs by photocatalysis

Pollutant Pollutantconc. (mM)

Catalyst Lamp CP t1/2(min)

kCP(moll1 min1)

pH Reference

2-CP 0.78 TiO2 (2.0gl1) 100 W medium pressure Hg 170 3.89 106 2.5 [61]200 3.27 106 4.5

80 1.01 105 7.1

48 1.25 105 9.548 2.80 105 12.5

2-CP 0.078 TiO2 ( 2 g l1) 15 W black flue fluorescent 20 2.10 106 3.0 [139]80 7.78 107 5.0

110 6.38 107 7.0140 4.51 107 9.0200 3.66 107 11.0

2-CP 0.78 TiO2 600 W high pressure xenon lamp 59 9.10 106 2.83.0 [43]0.54 53 7.71 106

0.39 36 7.47 106

0.19 27 5.04 106

0.077 15 3.72 106

2-CP 0.50 -Fe2O3 (1.5gl1) Solar simulator 420 2.57 106 6.0 [15]2-CP 1.00 PW12O403

(7.0 104 moll1)Oriel 1000 W Xe arc lamp 8 8.60 105 1.0 [7]

3-CP 8,8 8.20 105

4-CP 7,3 9.50 105

4-CP 1.00 TiO2 125 W high pressure Hg lamp 180 2.70 104 [105]TiO2/AC-M 160 6.40 104

TiO2/AC-PC 60 5.24 105

4-CP 1.00 TiO2 High pressure Hg lamp 125 W 256 2.70 104 [172]TiO2-ACM 108 6.40 104

TiO2-ACpc 256 2.70 104

2,3-DCP 0.50 -Fe2O3 (1.5gl1) Solar simulator 900 2.33 106 6.0 [15]2,4-DCP 0.12 TiO2 (0.1gl1) 125 W high pressure Hg lamp 22 4.03 106 6.0 [124]2,4-DCP 0.34 TiO2 (g l1) Solar light in flat reactor 845 2.82 107 6.6 [84]

0.59 0.1 1368 2.98 107 6.20.57 0.2 634 6.24 107 5.90.56 0.5 433 8.93 107 6.2

0.60 0.5 1061 3.93 107 7.60.59 1.0 499 8.17 107 5.90.58 2.0 416 9.59 107 5.60.55 0.1 Solar light in CPCs modules 252 1.50 106 4.60.55 0.1 354 1.07 106 5.90.53 0.2 173 2.11 106 4.20.56 0.5 145 2.66 106 4.20.58 0.5 528 7.56 107 5.90.56 1.0 234 1.65 106 6.00.55 2.0 238 1.61 106 5.1

2,4-DCP 0.50 -Fe2O3 (1.5gl1) Solar simulator 300 2.68 106 6.0 [15]2,4-DCP 1.00 PW12O403

(7.0 104 moll1)Oriel 1000 W Xe arc lamp 8,4 8.60 105 1.0 [7]

2,6-DCP 10,4 6.60 105

3,4-DCP 5,8 1.26 104

3,5-DCP 5,3 1.28 1043,5-DCP 0.30 TiO2 (0.1gl1) 125 W high pressure Hg lamp 165 9.32 107 7.0 [124]

2,3,5-TCP 0.43 TiO2 (0.1gl1) 125 W high pressure Hg lamp 59 2.32 106 7.0 [124]2,4,6-TCP 1.00 PW12O403

(7.0 104 moll1)Oriel 1000 W Xe arc lamp 10 7.50 105 1.0 [7]

2,4,6-TCP 0.50 -Fe2O3 (1.5gl1) Solar simulator Not reached 1.17 106 6.0 [15]

PCP 0.65 TiO2 (0.1gl1) 125 W high pressure Hg lamp 106 2.96 106 7.0 [124]


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[155,216]. One possibility of coupled conductors is basedon a combination of TiO2/CdS which allows higher degra-dation efficiency of 2-CP than the single semiconductorsystem [61]. The enhancement of photodegradation effi-ciency of 2-CP in TiO2/CdS/UV system may be attributedto the rapid transfer of the photogenerated electrons from

the cadmium sulfide to the titanium dioxide conductionband, resulting in the effective separation of the electronsand holes [216]. The addition of activated carbon to titaniahas been also studied and the results shows that the additionof a commercial H-type activated carbon to titania underUV radiation induces a beneficial effect on the photocat-alytic degradation of 4-CP whereas the addition of L-typecarbon is rather detrimental [172].

7.2.1. Effect of pH

The pH has a great effect on the photodegradation effi-ciency of CPs. The literature [182,183,222] suggests thatTiO2 surface carries a net positive charge at low pH value,

while the CPs and intermediates are primarily negativelyand neutrally charged. Therefore, low pH values can facil-itate the adsorption of the organic molecule and promotebetter photocatalytic degradation. This statement is in agree-ment with the results obtained by Ku et al. [139], wherethe photodegradation of 2-CP in the presence of TiO2 un-der various solution pH showed higher pseudo-first orderphotolytic rate constant at pH 3. However, better removalefficiency of CPs in TiO2/UV system under alkali condi-tion has been also reported [212,216,222]. One possibleexplanation is that the photocatalytic transformation of CPsdoes not involve hydroxyl radical oxidation exclusively

and direct electron transfer and surface sorption reactionsalso contribute significantly to the disappearance of CPs inTiO2 suspensions [212,222]. Another possible explanationis that TiO2 has an amphoteric character with a point zerocharge around pH 6 [274], and the substrate can undergoacidbasic equilibrium. Consequently, the adsorption of thesubstrate may be affected, strongly influencing the degrada-tion rate [137]. The results presented by Doong et al. [61]and Rideh et al. [203] show that the rate constant is lowerin acid medium than in basic medium for the photodegra-dation of 2-CP in the presence of TiO2. Tang and Huang[228] studied the photodegradation of 2,4-DCP and foundthat the optimal pH for this system was pH 5.

7.2.2. Effect of the dose of semiconductor

The influence of the dose of the semiconductor in thephotocatalytic reaction has been studied in the last years.At low photocatalyst loadings, the removal of organic com-pounds increased linearly with the catalyst loading; how-ever, the presence of excess photocatalyst in the aqueoussolutions could cause a shielding effect on the penetrationlight [84] [275]. After the achievement of the maximumreaction rate, increased turbidity of the solution reducedthe light transmission through the solution, while belowthis level, it is assumed that the catalyst surface and the

absorption light by TiO2 were limiting [203]. Doong et al.[61] observed that increasing the TiO2 dosage decreasedthe photodegradation efficiency of 2-CP. Removal of 71%was observed with the illumination time of 150 min when0.1gl1 was amended into the system. However, only 50%of 2-CP were decomposed when the dosage of TiO2 in-

creased to 1 g l1

. Rideh et al. [203] carried out a series ofexperiments to find out an optimum catalyst concentrationby varying the concentration of TiO2 on the reaction with2-CP from 1.2 102 to 0.3kgm3. The results showedthat the initial photodegradation rate increases linearly withcatalyst concentration up to 0.2 kg m3.

7.2.3. Effect of the initial concentration of CP

Many researchers have studied the effect of the initialconcentration of the pollutant in the photocatalysis. In allthe cases it was found that an increase in the concentrationof the contaminant involves a decrease in the photocatalyticreaction rate. A plausible explanation can be that since the

initial concentration increases, more and more organic sub-stances are adsorbed on the surface of TiO2, but the intensityof light and illumination time are constant; consequently,the OH formed on the surface of TiO2 is constant, the rel-ative number of OH attacking the CPs decreases, and thusthe photodegradation efficiency decreases too [178]. Chanet al. [43] and Rideh et al. [203] observed that an increase inthe initial concentration of 2-CP when degraded photocat-alytically by TiO2 involved a decrease in the rate of degra-dation. Chen et al. [46] studied the photocatalytic degra-dation of 2-CP, 2,4-DCP and 2,4,6-TCP in the presence ofmanganese ions at pH 3 varying the concentration of the

CPs. In all the cases, the rate constant of the photodegrada-tion decreases when increasing the initial concentration ofthe CPs.

7.2.4. Effect of temperature

Temperature has been proved to be an important aspectwhen studying the photocatalysis of CPs. Rideh et al. [203]and Chen et al. [46] studied the influence of temperature onthe rate of degradation of 2-CP. Rideh et al. [203] affirmthat the oxidation rate of 2-CP did not change significantlyin the range of 1565 C. Chen et al. [46] reported that therate increases following Eq. (67):

k = 0.0043T 1.2146 (min1, T in K) (68)

7.2.5. Effect of the oxygen concentration

Axelsson et al. [276], Barbeni et al. [16] and Chan et al.[43] have reported that the partial pressure of oxygen is acrucial factor in the photocatalytic reaction, in fact, photo-catalysis is commonly carried out with a source of oxygenor air [124,139,203,276278]. The limitation of the rate ofphotocatalytic degradation is attributed by most researchersto the recombination of photogenerated holeelectron pairs.Oxygen absorbed on the surface of TiO2 prevents the re-combination process by trapping electrons according to the


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reaction [203]:

O2 + e O2

Rideh et al. [203], found that the degradation rate of 2-CPincreases non-linearly with oxygen partial pressure. It wasassumed that the reaction rate is a function of the fraction ofadsorption sites occupied by oxygen; hence oxygen adsorp-tion becomes a limiting factor at very low dissolved oxygenconcentrations.

8. Conclusions

Taking into account only the half reaction time data be-tween the different AOPs studied for the degradation ofCPs (shown in Tables 57, 9, 11, 12 and 15) it can beconcluded that single ozonation appears to be the mostsuitable method for the treatment of aqueous solutionscontaining CPs. In fact, by means of single ozonation, the

lowest half-times have been reported for the degradation ofMCPs, TCPs and TTCPs. However, AOPs based on H2O2(Fenton, photo-Fenton and H2O2-UV) have been also re-ported to achieve the degradation of this compounds in ashort period of time. Moreover, these techniques are also ofinterest because they involve less cost expenses than ozona-tion. As for the degradation of TTCPs, besides ozonation,the combination O3/UV and photo-Fenton techniques havebeen reported to achieve high reaction kinetic rates. On theother hand, the photo-Fenton process seems to be a suit-able technique for the treatment of DCP, jointly with singleozonation and Fenton which have also shown reasonableresults for its degradation.

According to the literature, photocatalytic processes haveshown higher half reaction times than the rest of the AOPsfor the treatment of most of the CPs. However, photocatal-ysis displays also some advantages with respect to the otherAOPs, since no oxidant is consumed during the oxidationand it does not require any further separation after the reac-tion (like in the case of Fenton-like processes, where Fe2+

ions are dissolved in the medium during the reaction).

Acknowledgements

The authors wish to express their gratitude for the eco-nomic support given by the Spanish Ministry of Educa-tion and Culture (CICYT projects PPQ 2001-3046 and PPQ2002-00565). M. Pera-Titus and V. Garca-Molina are thank-ful to Prof. Juha Kallas from Lappeenranta University ofTechnology (Finland) and Dr. Sandra Contreras for the aidgiven in this work.

References

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Exposisicon Martes-13 - Marc Pera-AOPs

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