Synthetic endocrine disruptors in the environment and water

Journal of Environmental Management 85(2007)816–832

Review

Synthetic endocrine disruptors in the environment and water

remediation by advanced oxidation processes

Isil Gu ltekin,Nilsun H.Ince ?

Bogazici University,Institute of Environmental Sciences,34342Bebek,Istanbul,Turkey Received 19October 2006;received in revised form 19April 2007;accepted 24July 2007

Available online 4September 2007

Abstract

The present study is an overview of the literature on classes and types of compounds described as ‘‘endocrine disruptors’’and their treatability in water by advanced oxidation processes,which generate hydroxyl radicals in water.The review is limited to details of the destruction of three classes of endocrine disruptors,namely bisphenols,alkylphenols and phthalates,which are among the most highly suspected endocrine disrupting compounds that interfere with the hormonal system of wildlife.It was found that photocatalysis with titanium dioxide was the most frequently tested advanced oxidation method most likely due its potential to render complete mineralization.There is suf?cient research also with direct and indirect photolysis and ozonation,which were less effective for the overall mineralization but more representative of the conditions existing in real water treatment plants.r 2007Elsevier Ltd.All rights reserved.

Keywords:Endocrine disruptors (EDCs);Advanced oxidation processes (AOPs);Bisphenols;Bisphenol A;Alkylphenols;Nonylphenol;Phthalates;Byproduct analysis;Hydroxyl radical

Contents 1.Introduction ...............................................................................8172.

Overview of industrial endocrine disruptors .........................................................8172.1.Bisphenols ............................................................................8182.2.Alkylphenols ..........................................................................8182.3.Phthalates ............................................................................8183.

Advanced oxidation processes (AOPs)as potential methods of EDC destruction in the water environment ............8183.1.Destruction of bisphenol A by advanced oxidation processes ........................................

8193.1.1.Direct and indirect photolysis with UV..................................................8193.1.2.Photocatalysis with TiO 2............................................................8203.1.3.Dark advanced oxidation reactions with ozone,ultrasound and electrochemical processes..............8213.1.4.Overview of reaction conditions and oxidation byproducts....................................8223.2.Destruction of nonylphenol by advanced oxidation processes ........................................

8233.2.1.Direct and indirect photolysis with UV..................................................8253.2.2.Photocatalysis ...................................................................8253.2.3.Dark advanced oxidation reactions ....................................................8263.2.4.Overview of reaction conditions and oxidation byproducts....................................8263.3.Destruction of phthalates by advanced oxidation processes..........................................

8263.3.1.Direct and indirect photolysis with UV..................................................

826

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?Corresponding author.Tel.:+902123597038;fax:+902122575033.

E-mail address:ince@https://www.wendangku.net/doc/9f5646566.html,.tr (N.H.Ince).

3.3.2.Photocatalysis with TiO2 (827)

3.3.3.Dark advanced oxidation reactions (828)

3.3.4.Overview of reaction conditions and oxidation byproducts (828)

4.Conclusions (829)

Acknowledgments (830)

References (830)

1.Introduction

Disruption of the endocrine system in living organisms by synthetic organic chemicals has been of great concern in recent years due to the recognition that the environment is contaminated with numerous‘‘endocrine disrupting com-pounds’’(EDCs)that exert hormonal activity(Metzler and Pfeiffer,2001).An endocrine disruptor by de?nition is an exogenous agent that interferes with the synthesis, secretion,transport,binding,action and elimination of natural hormones in the body,which are responsible for the maintenance of homeostasis,reproduction,develop-ment and behavior(US Environmental Protection Agency (US EPA),2001).

The primary function of an endocrine system is to transform various exogenous stimuli into chemical mes-sengers and hormones,resulting in appropriate gene expressions and synthesis of proteins and/or activation of already existing tissue-speci?c enzyme systems(Jintelmann et al.,2003).The three major endocrine disruption endpoints are estrogenic(compounds that mimic or block natural estrogens),androgenic(compounds that mimic or block natural testosterone),and thyroidal(compounds with direct and/or indirect impacts on the thyroid)(Synder et al.,2003).EDCs are linked to a variety of adverse health effects in wildlife,such as hormone-dependent cancers, reproductive system disorders and reduction in reproduc-tive?tness(Fry et al.,1987;Synder et al.,2003;Alum et al., 2004).Endocrine disruptors may be natural or more commonly synthetic,all with very diverse chemical structures(Metzler and Pfeiffer,2001).The most promi-nent classes of chemicals that contain EDCs are natural estrogens,phytoestrogens,pesticides(methoxychlor),sur-factants(nonylphenol),plasticizers(diethylphthalate, BPA)and organohalogens(PCBs and dioxin)(US EPA, 2001).

The majority of EDCs are ubiquitous as they may be present in all compartments of the environment(water,air, soil and sediments)upon imperfect manufacturing pro-cesses and/or leaching from?nal end products.Sources of surface water contamination with EDCs are sewage ef?uents from domestic and industrial facilities and industrial ef?uent discharges(Gomes and Lester,2003). The major source of EDCs in domestic sewage is the daily produced male and female hormones and/or ingested synthetic steroids,which are excreted with urine and discharged into municipal systems,where they may be partially removed by biochemical oxidation before being released to surface waters as active estrogens(Deborde et al.,2005;D’Ascenzo et al.,2003;Panter et al.,1999; Ternes et al.,1999).The concentration of EDCs in surface waters is variable.Ranges of ng là1are reported in some sites for natural synthetic hormones(Ying et al.,2002a; Belfroid et al.,1999)and m g là1for alkylphenols(AP)and bisphenols(Blackburn and Waldock,1995;Ahel et al., 1994).Such levels were found to induce feminization in male?sh and intersex induction(Gray and Metcalfe,1997). Although some researchers have reported a relation between reproductive abnormalities in animals and ex-posure to EDCs in real cases(Guillette et al.,1994; Andersson et al.,1988;Morrison et al.,1985),there is no apparent correlation established so far between EDC exposure and reproductive disorders in humans except for published reports during the last decade on the increase in estrogen-dependent cancers and the reduction in human sperm production(Herbst et al.,1989;Mocarelli et al., 1996).The risk for groundwater contamination arises from in?ltration of rainwater through land?lls and percolation from agricultural areas(Gomes and Lester,2003).Despite their low concentration in the aquatic environment,the EU has labeled hormone-like chemicals as‘‘hazardous’’due to the fact that even a trace amount of them is suf?cient to initiate estrogenic activity.

Conventional water and wastewater treatment plants are inef?cient for substantially removing many EDCs.For this reason,it is essential that future research focus on the investigation of appropriate treatment methods that can be integrated into water and wastewater treatment facilities to prevent the release of EDCs into the natural waters.The research with advanced oxidation processes(AOPs)during the last decade has shown that they are promising techniques for the removal of refractory pollutants from ef?uents of bio-treatment plants and/or surface waters. Hence,the aim of this study was to review and compare published research on the degradability of common endocrine disruptors,namely bisphenol A(BPA),non-ylphenol and phthalates by light and dark advanced oxidation techniques and to point out if available the ef?cacy of the studied method for the degradation and estrogenic activity removal of the parent compound.

2.Overview of industrial endocrine disruptors

Among numerous EDCs used in industrial processes,the three major classes that have recently received scienti?c and public interest as the most potential disruptors are bisphenols,AP and phthalates.The common feature of

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all is that they are produced in massive quantities,and a substantial fraction is released into the environment.

2.1.Bisphenols

The prototype of bisphenols is BPA,which is used mainly as a monomer for the production of polycarbonate, epoxy and unsaturated polyester-styrene resins,?ame retardants,fungicides,antioxidants and rubber chemicals. In addition,BPA is consumed as a resin in dental?llings, as coatings on cans,as powder paints and as additives in thermal paper(Fromme et al.,2002;Staples et al.,1998). Some researchers have shown that BPA could stimulate proliferation and synthesis of progesterone receptors of human breast cancer MCF-7cells,displace estradiol from estrogen receptors,induce synthesis of vitellogenin in cultivated trout liver cells and provoke transcription of recombinant yeast cells(Perez et al.,1998;Pawlowski et al., 2000;Routledge and Sumpter,1996).The estrogenic potency in vitro is10à3to10à4relative to that of estradiol, and4?10à4to10à5in vivo relative to that of diethyl-stilbestrol(Jintelmann et al.,2003).

BPA is released into the aquatic environment from industrial discharges,land?ll leachate and water streams containing plastic debris(Yamamoto and Yasuhara,1999; Yamamoto et al.,2001).The concentration range shows regional variations,not exceeding parts per billion levels in natural waters,but reaching much higher values in streams contaminated with industrial discharges.

2.2.Alkylphenols

AP are used as antioxidants and plasticisizers,and in the production of alkylphenol polyethoxylates(APEO)—the most widely used non-ionic surfactants(that comprise6% of total surfactant production worldwide)(Metzler and Pfeiffer,2001,Nimrod and Benson,1996).Of all the APEO production,about80%is released to the market as nonylphenol polyethoxylates(NPEOs)(Ying et al., 2002b),which are widely used in the production of plastics, textiles and agricultural chemicals,and in household goods such as detergents,paints,pesticides and cosmetics(Naylor et al.,1992;Nimrod and Benson,1996).Accordingly,a large majority of the research related to APEO and AP destruction in the water environment is focused on nonylphenols(NP),which enter the environment through wastewater streams with NPEOs.Under anaerobic condi-tions such as those found in sewers,sediments,and bio-treatment operations at wastewater treatment plants, NPEO is oxidized to NP,which is recognized with its stability,aquatic toxicity and estrogenic activity(Comber et al.,1993;McLeese et al.,1981;Soto et al.,1995). Research has shown that the degradation products of many APEO surfactants that are discharged with waste-water treatment plant ef?uents are estrogenic(Routledge and Sumpter,1996).2.3.Phthalates

Phthalates represent a class of chemical compounds used most widely as plasticizers for polyvinyl chloride(PVC) resins and cellulose?lm coating.To a minor extent they have also found application in cosmetics,insect repellents and propellants(Vitali et al.,1997).There are about60 different phthalates produced and consumed worldwide for very diverse purposes.The most important one in terms of production and consumption is2-ethylhexyl phthalate (DEHP),which is used mainly in improving technical properties of plastic materials such as PVC,and to a minor extent in the production of cosmetics,adhesives,paints and other daily used chemicals(Metzler and Pfeiffer,2001). Other prominent phthalates are butyl benzyl phthalate (BBP)and dibutyl phthalate(DBP),which have found application in vinyl?oor tiles and plasticizers,respectively (Metzler and Pfeiffer,2001).Phthalates are easily trans-ported to the environment during manufacture,disposal and leaching from plastic materials,in which they are bonded non-covalently to allow the required degree of ?exibility(Nilsson,1994).The short-chained phthalates, dimethyl phthalate(DMP)and diethyl phthalate(DEP)are among the most frequently identi?ed phthalates in diverse environmental samples including surface marine waters (Tan,1995;Fatoki and Vernon,1990),freshwaters(Vitali et al.,1997;Staples et al.,2000)and sediments(Tan,1995; Thuren,1986).Epidemiological studies with humans have shown that phthalates induce adverse health effects such as disorders in male reproductive tract,breast and testicular cancers and neuroendocrine system disruption(Spelsberg and Riggs,1987;Sharpe and Shakkebaek,1993).

3.Advanced oxidation processes(AOPs)as potential methods of EDC destruction in the water environment AOPs are those processes that involve in situ generation of highly reactive species such as the hydroxyl radical, which is the most powerful oxidizing species after?uorine with an oxidation potential of 2.80V(Parsons and Williams,2004).Unlike many other radicals,hydroxyl radical is non-selective and thus readily attacks a large group of organic chemicals to convert them to less complex and less harmful intermediate products.At suf?cient contact time and proper operation conditions,it is practically possible to mineralize the target pollutant to CO2,which is the most stable end product of chemical oxidation.The remarkable advantage of AOPs over all chemical and biological processes is that they are totally ‘‘environmental-friendly’’as they neither transfer pollu-tants from one phase to the other(as in chemical precipitation,adsorption and volatilization)nor produce massive amounts of hazardous sludge(as in activated sludge processes)(Ince and Apikyan,2000).

The most common AOPs developed for water and wastewater remediation are presented in Table1.Some of these processes such as photolysis with more than3000

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applications in Europe(as a disinfection process)and a large number in the US(for treating groundwater pollutants)are commercially available(Parsons and Williams,2004).Other processes such as combinations of H2O2,O3and UV,Fenton’s reagent,super-critical water oxidation and ionizing radiation have all been used at full scale.Photocatalysis and ultrasound have been assessed only at laboratory bench and pilot scales.It should be noted that research on the application of AOPs in dual or triple combinations of individual processes offer signi?cant kinetic and performance advantages.

Methods of AOP employed in studying the degradability of endocrine disruptors in water are:(i)direct(with UV) and indirect photolysis(with UV/O3,UV/H2O2,UV/ Fenton,UV/Fenton/oxalate);(ii)photocatalysis with TiO2or other semiconductors;and(iii)dark advanced oxidation reactions with ozone,electrochemical(EC) processes and ultrasonic cavitation.

3.1.Destruction of bisphenol A by advanced oxidation processes

3.1.1.Direct and indirect photolysis with UV

UV photolysis has been one of the most widely investigated advanced oxidation methods of BPA destruc-tion based on the facts that UV irradiation is a common practice of drinking water treatment,and photolysis is the main abiotic degradation pathway of organic matter in natural waters.The following is an overview and evalua-tion of such research carried out for destroying BPA. Rosenfeldt and Linden(2004)studied the degradability of BPA by low and medium pressure mercury UV lamps and reported5%and10–25%reduction by direct photolysis,respectively.However,exposure of the same samples to UV/H2O2at a?uence rate of1000mJ cmà2 (typical of the irradiation rate in water treatment plants) resulted in more than90%BPA destruction regardless of the UV source.They also recorded nearly100%reduction in estrogenic activity(EST)by UV/H2O2application.

A similar study reported by Chen et al.(2006)has shown that a low pressure UV lamp even when operated at 5000mJ cmà2is ineffective alone,while the ef?cacy is improved to80%and78%for BPA and EST removal, respectively,in the presence of10mg là1H2O2.The important operation parameter in the UV/H2O2process is the concentration of H2O2,which should be maintained high enough to ensure suf?cient OH radicals in solution and low enough to prevent the reaction of excess H2O2 with OH radicals.The impact of H2O2concentration as reported by Chen et al.(2006)was such that while97% BPA and EST removal was obtained at25mg là1H2O2,no further enhancement was obtained when the concentration was doubled.

The composition of the water containing the contami-nant is a crucial parameter of system ef?ciency in AOP. In accordance,Neamtu and Frimmel(2006a)reported that the degradation and reduction in EST activity of BPA by254nm UV irradiation was more ef?cient in puri?ed or surface water than in sewage water due to the presence of competing substances in the latter.They also observed a decline of EST in both matrices with increased irradiation time.In another study related to the impact of water composition,Zhan et al.(2006)reported that the degradation of BPA in the presence of natural humic substances by solar radiation was faster than in pure water.Consequently,they suggested a mechanism to the decay process that involved the:(i)production of excited BPA molecules leading to direct photolysis;and (ii)production of hydroxyl radicals from photoreactive components of humic substances leading to indirect photolysis.The impact of water matrix was further investigated by Zhou et al.(2004)in synthetic samples containing BPA and ferric oxalate(Fe-Ox)complexes exposed to a high pressure Hg lamp.Based on the phenomenon that Fe-Ox complexes in atmospheric waters produce H2O2,which react with Fe(II)to generate OH radicals according to Fenton reaction(Zepp et al.,1992), they found that maximum ef?ciency in their system was obtained at pH?3.5in a Fe/Ox molar ratio of10:120. They also reported that total mineralization(160min exposure)in these conditions was24%as opposed to 7.3%mineralization in the absence of oxalate.

The potential of Fenton reaction in UV photolysis of BPA was also studied by Katsumata et al.(2004)using a Xe lamp emitting UV light at l o300nm.The study showed that maximum destruction could be accomplished at pH?3.5–4.0with a H2O2–Fe(II)ratio of10(by M),at which50%mineralization was possible in24h.The optimal ratio of H2O2/Fe(II)/BPA for complete degrada-tion of BPA was reported as9:0.9:1.

Finally,indirect photolysis of BPA was investigated by Irmak et al.(2005)via an O3/UV technique using a low pressure Hg UV lamp.They observed that at a dose of 18.7?10à3mmol minà1O3,BPA was completely con-verted consuming1.40and1.49mmol O3in the presence and absence of UV irradiation during75and80min exposure,respectively.

Table1

Most common AOPs evaluated for water and wastewater treatment

Advanced oxidation processes

Photochemical processes Non-photochemical processes

UV oxidation processes Ozonation

UV/H2O2Fenton

UV/O3Ultrasound(US)

UV/H2O2/O3US/H2O2,US/O3,US/Fenton

UV/Ultrasound Electrochemical oxidation

Photo-Fenton Supercritical water oxidation

Photocatalysis Ionizing radiation

Sonophotocatalysis Electron-beam irradiation

Vacuum UV(VUV)Wet-air oxidation

Microwave Pulsed plasma

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3.1.2.Photocatalysis with TiO2

Photocatalytic reactions via photo-generated holes of TiO2have been thoroughly investigated for water remedia-tion during the last couple of decades.Under near UV irradiation,TiO2is photoactivated to form active oxygen species(e.g.OH radicals)on surfaces of the crystal,and the radicals readily react with a wide range of organic moieties to ultimately produce CO2.The apparent advantage of this method over homogeneous AOPs discussed above is its outstanding potential to render complete mineralization. Studies reported in the literature on photo-decomposition of BPA in the presence of TiO2involve the use of powder, sheet or composite TiO2?lms and comparative assessment of the degree of BPA decay,EST reduction and miner-alization.A number of studies are devoted to modi?cation and optimization of the system via variations such as replacement of the arti?cial light source by solar irradia-tion;introduction of rotating or composite TiO2-metal sheets,addition of chemical reagents;deposition of metals and/or reticulated carbon;and hybridization of the photo-reactor with low-pressure submerged membranes.The following discussion covers a critical review of the literature on TiO2-catalyzed photolytic decomposition of BPA. Ohko et al.(2001)reported the degradation of BPA in TiO2aqueous suspensions irradiated by a Xe lamp.They found that total disappearance and mineralization required 15and20h,respectively,and all intermediate products vanished within that time.They also reported that transcriptional estrogenic activity in response to human estrogen receptor was reduced to less than1%of the initial activity within4h.The reaction pathway for the process as proposed by Watanabe et al.(2003)followed an initial attack of OH and OOH radicals on the two methyl groups and the subsequent cleavage of the methyl moieties to produce simple aldehydes,acids and carbon dioxide.In an attempt to improve the ef?ciency of the process,Wang et al.(2006)added b-cyclodextrin (b-CD)and monitored the degree of decomposition and mineralization of BPA.Under optimum conditions (b-CD? 4.4?10à5M,pH?6),the rate of decomposition was accelerated by23%upon b-CD addition,while complete and37%mineralization was obtained,respec-tively,in the presence and absence of b-CD during a total contact of120min.The improvement was attributed to enhanced adsorption of BPA on TiO2surfaces and weakening of the bond energy between atoms of BPA molecule via‘‘inclusion interactions’’with b-CD.Another study aimed to improve the ef?ciency of BPA decomposi-tion in illuminated suspensions of TiO2was conducted by Horikoshi et al.(2004)in a plain,thermal and microwave-assisted non-thermal hybrid system.It was found that while 67%mineralization was accomplished in the plain reactor in90min,the ef?ciency was increased to90%in the hybrid reactor with either microwave irradiation or external heating.Different reaction byproducts in thermal and microwave irradiation routes were attributed to adsorption mode variations of BPA on TiO2particle surface.

Enhancement of BPA decomposition in suspensions of TiO2was further studied by Xie and Li(2006)using(i)a TiO2/Ti?lm electrode;(ii)a gold-deposited TiO2?lm;and (iii)a reticulated vitreous carbon(RVC)electrode that continuously generated H2O2.The study showed that gold deposition enhanced photocatalytic and photoelectrocata-lytic activities under both UV and visible irradiation by improving the ef?ciency of eà/h+separation on the conduction band of TiO2.Furthermore,the use of RVC cathode was found to signi?cantly assist the photoelec-trocatalytic reaction and improve the degree of BPA degradation.

Chiang et al.(2004)compared the ef?ciency of plain and platinized TiO2on photolytic decomposition and miner-alization of BPA at pH3and10.They found that at optimized Pt loading,the photocatalytic activity of the solution could be enhanced by3–6times at both pH.The relatively low degree of mineralization at pH10was attributed to the negative charge on TiO2surfaces(that inhibits adsorption)and the presence of carbonate ions (that scavenge OH radicals)formed during oxidation of organic carbon.On the other hand,Coleman et al.(2005) reported that enhancement of photocatalytic activity by immobilization of TiO2on Pt or Ag was dependent on the concentration of BPA and that platinization was ineffective in weakly concentrated solutions,typical of natural waters. They reported that platinum accelerated the rate of mineralization only if the concentration of BPA was larger than50mg là1.In a similar attempt to enhance the overall degree of decomposition in TiO2catalytic process,Naka-shima et al.(2002)studied the photocatalytic decay of BPA in the presence of immobilized TiO2particles on still and rotating polytetra?uoroethylene(PTFE)mesh sheets illu-minated by black?uorescence.They found that the rate of decomposition was twice faster in the presence of rotating sheets as a consequence of enhanced mass transfer of BPA to TiO2surfaces upon rotational motion.Finally,Lee et al. (2004)investigated the ef?ciency of BPA destruction in an immobilized TiO2?lm prepared by a sol–gel method and located on the external surface of a glass tube.Their results showed that highest BPA removal was achieved when titanium–sol solution was used as a binder.The optimum coating number and the?lm thickness for maximum yield was3and5.29m m,respectively.While68%BPA removal was observed with single coating,the fraction increased to 98%in the presence of3-fold coating,above which the ef?ciency was reduced by increased competition for UV radiation.

A study devoted to con?guration of the photoreactor for practical applications in water puri?cation was conducted by Zhang et al.(2003)in a tubular ceramic reactor,with inner surface coated with Pt-loaded TiO2?lm and irradiated with a UV lamp located longitudinally in the center.The water contaminated with BPA was circulated through the reactor during air injection.It was found that the proposed reactor had higher photocatalytic activity in the presence of optimal aeration conditions

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than conventional reactors and100%BPA removal was possible in4h during aeration at1dm3minà1.

The feasibility of photocatalytic systems using arti?cial light sources was questioned by Kaneco et al.(2004),who investigated the impact of natural solar radiation as an alternative to the costly and energy-intensive arti?cial sources, and identi?ed the optimal values of the catalyst dose,initial BPA concentration,temperature and light intensity.It was found that TiO2-catalyzed processes aimed to destroy BPA in aqueous media were most economic at C0(BPA)?20mg là1, TiO2?10mg là1,T?401C,pH? 6.0,and solar light intensity?0.35mW cmà2.The authors reported complete mineralization in11h at these optimal conditions. Although photocatalysis by TiO2seems to be an effective method for the overall degradation of BPA,some researchers were concerned about the practicality of the process related to the relative dif?culty of separating the catalyst powders and/or crystals from solution at the end of treatment.A study devoted to this problem was conducted by Fukahori et al.(2003),who attempted to modify TiO2crystals by combining them with a strong adsorbent-zeolite to form a paper-like composite.They found that without UV irradiation,TiO2–zeolite composite sheets were much more ef?cient in BPA removal than free TiO2sheets and powders,whereas in the presence of black light,powders were more effective.By contrast,miner-alization was highest in the presence of illuminated TiO2–zeolite sheets,which the authors attributed to the adsorptive removal of the oxidation intermediates onto zeolite surfaces.In conclusion,composite sheets were strongly recommended for removing BPA from water due to:(i)their larger ef?ciency than ordinary TiO2sheets for mineralization,(ii)their potential to capture the inter-mediate byproducts,preventing release into water;and (iii)their unproblematic separation from solution.

A similar attempt to ease the separation and recovery of sub-micron sized TiO2catalysts from the treated solution was conducted by Chin et al.(2007)in a hybrid system combining photocatalysis and membrane?ltration in a single module.The system consisted of a low-pressure submerged hollow?ber membrane reactor in direct contact with TiO2and BPA and irradiated by black light ?uorescent lamps.Aeration was applied to reduce the fouling of the submerged membrane and to keep the medium in suspension.It was found that97%of BPA was degraded after70min,and90%of it was mineralized after 100min contact with the hybrid system.An intermittent frequency(ratio of off-time to on-time)of0.1was reported as the optimum value to stabilize the sustainability of the membrane,and an aeration rate of0.5l minà1was the optimum for providing excess dissolved oxygen and for improving the mass transfer process.

3.1.3.Dark advanced oxidation reactions with ozone, ultrasound and electrochemical processes

Ozonation has been one of the most widely investigated techniques of advanced oxidation owing to the fact that it is commonly used in a large number of water treatment plants as a clarifying and disinfecting agent.The mechan-ism of organic matter removal in ozonated waters is either direct oxidation by molecular ozone or indirect oxidation by OH radicals that are formed by the decomposition of ozone in alkaline conditions.In natural waters,a large fraction of these radicals(which are much less selective and much more reactive than ozone)are scavenged by the water matrix(von Gunten,2003); therefore the principle pathway of oxidation is direct reaction of ozone with the contaminant.The literature on BPA destruction by ozone is focused on the investigation of reaction kinetics,estrogenic activity and the operation parameters such as ozone feed rate and pH,which dictates the distribution of molecular and radical species.Bypro-duct analysis of dark ozone processes in water containing BPA has not been published so far,but is under investigation in our laboratory.The following discussion covers a critical review of published research on ozone-based destruction of BPA in water.

Lee et al.(2003)studied the degradation of BPA at pH?2,7and12and found that the rate was sensitive and proportional to the ozone feed rate,but insensitive to pH and the concentration of BPA.They found that even at extreme alkaline conditions(e.g.pH?12),where OH radicals are expected to dominate over O3,the rate of degradation remained unchanged,as a consequence of the mass transfer resistances,the deprotanation of BPA,and the competition of d OH with ozone.Moreover,the addition of H2O2was found to suppress the reaction by the competition between BPA and H2O2for O3and OH radicals.The direct rate constant of BPA–O3reaction at pH2and12was estimated as 1.3?104Mà1sà1and 1.6?109Mà1sà1,respectively.In another study devoted to kinetics of BPA decomposition by ozone in the presence of a strong d OH scavenger,Deborde et al.(2005)found that the apparent rate was minimum at pH o5and maximum at pH410.The reactivity of ozone with BPA and its two ionized species was reported as1.68?104,1.06?109and 1.11?109Mà1sà1,respectively.

Some researchers were concerned with the practicality of ozonation process in real water treatment plants.A study related to this problem was conducted by Alum et al. (2004)by monitoring the concentration and estrogenic activity of BPA during120min ozonation at conditions similar to those of drinking water treatment plants.They found that99%of BPA was converted with an initial ozone concentration of30m M in less than2s.However,the level of estrogenicity increased during initial contact, declined after subsequent contact and stabilized after 60min,signifying that the oxidation byproducts had higher estrogenic activity than BPA and some residual estrogenic response remains after treatment.Behavior of BPA in oxidative water treatment facilities was also studied by Lenz et al.(2004),who reported1.4mg là1as the optimum O3concentration for complete BPA destruction and5min as the suf?cient contact time for removing the endocrine

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activity without the formation of adsorbable organic halogens.

On the other hand,Irmak et al.(2005)showed that in laboratory conditions with a much higher BPA concentra-tion and continual ozone injection,ozonation was more ef?cient in removing17b-estradiol(which is used as an indicator of estrogenic activity)from water than removing BPA.They found that the stoichiometric molar ratio and contact time for complete conversion of BPA(0.40mM)by ozone were14.94and80min,respectively,and the ratio was 1.68times larger than that required for complete removal of 17b-estradiol.Disagreement in the changes observed in estrogenic or endocrine activities in different laboratories is due to variations in the applied assay or technique and differences in the method of response evaluation. Finally,Kamiya et al.(2005)investigated the impact of competing organic compounds in sewage ef?uents on the ef?ciency of BPA destruction,EST activity and toxicity reduction by ozone.They found that BPA could be successfully decomposed by ozonation with the typical dose(3–5mg là1)regardless of the presence of other organic substances.They also reported that EST activity could be declined to below detection limit even with 1mg là1ozone,which was suf?cient to reduce the aquatic acute toxicity in real sewage as well.

The literature on BPA removal from water by EC processes,which carry out oxidation and reduction reactions on surfaces of electrodes at electrode–electrolyte interface,is limited.Some of these processes are‘‘direct’’applied without the addition of chemical reagents and some are‘‘indirect’’involving the addition of chemicals and generation of highly powerful oxidizing species.The study of Korshin et al.(2006)is about a direct EC process (in0.01M NaCl electrolyte)using an undivided?ow-through cell with a PbO2anode and a stainless-steel cathode.The ef?ciency of the process was tested by monitoring the concentration of chloroacetic acids,TOC and adsorbable organic halogens(TOX)in the ef?uent. The study showed that under the experimental conditions employed,EC-treatment of BPA resulted in the formation of identi?ed and non-identi?ed chlorinated byproducts. Mono-chloroacetic acid was reported as the primary identi?ed product(5.9%),while the non-identi?ed part of TOX was comprised of aromatic chlorinated forms of BPA,which were resistant to degradation by EC and would likely resist similar conditions in drinking water treatment plants.In the study of Jiang et al.(2005),a direct electrochemical oxidation process using graphite felt–tita-nium and titanium–graphite felt–stainless steel as the electrodes was tested on crude and bio-treated wastewater to investigate its ef?ciency for BPA destruction.They reported that while bio-treatment with post-sedimentation could lower BPA concentration from several hundreds to a hundred ng là1,levels as low as1–100ng là1could be achieved with electrochemical treatment.

An indirect EC process using the Fenton reaction as the source of hydroxyl radicals and involving the generation of Fenton’s reagent was tested by Go zmen et al.(2003).It was found that conversion of BPA increased with increasing [Fe2+]/[BPA]molar ratio,and replacement of ferrous/ ferric pair with cuprous/cupric pair resulted in faster BPA degradation and faster mineralization.However,ferrous/ ferric pair was more ef?cient in terms of conversion and total mineralization versus the charge utilized.

The use of ultrasonic pressure waves in water to generate hydroxyl radicals is a novel technique,which promotes the formation,growth and violent implosion of cavitation bubbles with the outcome of extreme local conditions (5000K,2000atm)and high energy chemistry(Suslick, 1990;Ince et al.,2001).The presence of impurities in water such as solids and dissolved gases largely enhances the ef?ciency of chemical reactions as a consequence of reduced cavitational threshold(Mason,1999).The litera-ture on ultrasonic means of BPA destruction is limited to a couple of investigations.Kitajima et al.(2006)studied the effect of dissolved gases during sonication of BPA samples at500kHz and found that the rate of degradation increased in the order O24Ar4air4N2.Ioan et al. (2007)compared Fenton and sono-Fenton processes in an ultrasonic bath operated at43–47kHz.They found that the degradation of BPA by the Fenton process was much faster in the presence of ultrasound than in its absence,and the rate was further enhanced with increasing concentra-tions of Fe(II)and decreasing levels of pH.

3.1.

4.Overview of reaction conditions and oxidation byproducts

Throughout the review,we found that differences in the applied experimental conditions in different studies are the principle reason for the variations in?nal outputs and conclusions.Although descriptive experimental features of the articles discussed in Sections3.1.1–3.1.3have already been speci?ed in the text,a comparative list of all is given in Table2.

One of the major concerns of BPA destruction by AOP is the?nal product of oxidation,because in some cases the product might still exert endocrine activity and/or might be more toxic than the parent compound.As a consequence, some of the published research was coupled with bypro-duct analysis,while others only with endocrine or estrogenic activity testing via bioassay techniques.In general,analysis of estrogenic activity showed that the ef?uent either exhibited no hormonal activity or much less that of the original sample.

A list of oxidation byproducts that were identi?ed in studies discussed so far is given in Table3.The list shows that phenol and p-hydroquinone were the two most commonly observed products regardless of the AOP technique.Hydroxyacetophenone was identi?ed in photo-Fenton and photocatalytic processes,while methylbenzo-furan was observed in photo-Fenton,some TiO2and ozonation processes.It was proposed that the initial step in the reaction mechanism of photocatalytic processes was adsorption of two OH moieties or BPA itself on TiO2

I.Gu¨ltekin,N.H.Ince/Journal of Environmental Management85(2007)816–832 822

surface,followed by d OH attack on the two phenyl carbons to produce short and long chain structures leading ?nally to oxalic,formic and acetic acids(Watanabe et al., 2003;Horikoshi et al.,2004;Wang et al.,2006).The proposed reaction mechanisms for homogeneous processes involved initiation either by direct attack of d OH to BPA, or hydroxylation followed by abstraction of a hydrogen atom from phenolic hydroxyl groups to form quinone-like compounds,which by further oxidation ended up in aliphatic acids(Fukahori et al.,2003;Kaneco et al., 2004;Katsumata et al.,2004).

3.2.Destruction of nonylphenol by advanced oxidation processes

Published literature on the removal of NP from water by AOP is limited,focusing on their degradability by indirect photolysis,photocatalytic oxidation,ozonation

Table2

Summary of reaction conditions for BPA removal from water by AOP

Concentration Reaction conditions References

Direct and indirect photolysis with UV

23.3m M1kW MP a or four15W LP b Hg lamps;H2O2?0–25ppm;UV?uence?0–1500mJ cmà2Rosenfeldt and

Linden(2004)

60m M15W LP Hg lamp(253.7nm);H2O2?0–50ppm;UV?uence?100–5000mJ cmà2;pH?5.3Chen et al.(2006) 520m M15W LP Hg lamp(254nm);H2O2:0–750m M;pH?6.7;photonic?ux?4.25?10à6Einstein sà1Neamtu and Frimmel

(2006a)

44m M500W MP Hg vapor lamp(365nm);I c?0.525mW cmà2;humic acids?10mg là1Zhan et al.(2006)

8.8–44m M125W HP d Hg lamp(X365nm);pH?3–8,[Fe(III)]:[Ox]?10/30,10/60,10/120(m M/m M)Zhou et al.(2004)

44m M990W Xe lamp(o300nm),I?0.5mW cmà2;pH?2–4.5;Fe(II)?0–4?10à5M;H2O2?0–4?10à4M Katsumata et al.

(2004)

400m M15W LP Hg lamp;O3fed at10.3–18.7mmol minà1;pH?5.25Irmak et al.(2005) Photocatalytic processes with TiO2

175m M200W Hg–Xe lamp(365nm),I?10mW cmà2;1g là1TiO2Ohko et al.(2001) 100m M75W Hg lamp(360nm),I?2.5mW cmà2;pH?4.4;TiO2?2g là1Watanabe et al.(2003) 22–88m M250W metal halide lamp(X365nm);1g là1TiO2;b-CD e?0–17.6?10à5M;pH2–12Wang et al.(2006) 100m M250W Hg lamp,I?0.9mW cmà2;TiO2?2g là1;pH?6.7;MW f power?300W Horikoshi et al.(2004) 49.3m M8W MP Hg UV lamp(365nm),I?0.68mW cmà2,110W HP Na vapor lamp(450–650nm),

intensity?49mW cmà2;Anode:TiO2/Ti or Au–TiO2/Ti;cathode:RVC or Pt;pH?6.17

Xie and Li(2006)

88m M20W black?uorescent lamp(355nm);0.1g là1platinized TiO2;TiO2:Degussa P25,Hombikat UV100,Milennium

PC50;pH?3,10

Chiang et al.(2004) 5.6m M

224.4m M

15W black?uorescent lamp(350nm);TiO2?1g là1,bare and coated with silver or platinum;pH?3Coleman et al.(2005) 0.396m M15W black?uorescent lamps(x2);I?0.24mW cmà2;bare and TiO2-modi?ed PTFE mesh sheets Nakashima et al.

(2002)

44m M6W black light blue lamp(365nm);aeration?3l minà1;immobilized TiO2particles;pH?4.5;T?301C Lee et al.(2004)

20.24m M30W Hg lamp(254nm);ceramic tube(inner surface coated with Pt-loaded TiO2and water–glass);solution

?ow?1.25m3hà1;aeration?1l minà1

Zhang et al.(2003)

0–440m M Solar illumination,I?0–1.7mW cmà2;TiO2?0–20g là1;pH?2–10;T?10–701C.Kaneco et al.(2004) 100m M Hg–Xe lamp(365nm);I?2mW cmà2;TiO2powder or TiO2–zeolite sheet(2g là1)Fukahori et al.(2003) 44m M8W black light?uorescent lamps(x4);TiO2?0.5g là1,Flux?100l mà2hà1;aeration?0.2–4l minà1;pH?4Chin et al.(2007) Dark advanced oxidation with ozone,electrochemical and ultrasonic processes

4.4–44m M O3feed?0.4–2mg minà1;pH?2,7,12;H2O2?1–10mM(added batch-wise or continuously)Lee et al.(2003)

0.1m M O3generated from pure O2;ozonation from a liquid O3stock solution(40mg là1)Alum et al.(2004) 400m M O3generated from oxygen,fed at10.3–18.7mmol minà1;pH?5.25Irmak et al.(2005) 0.002m M

0.44m M

O3generated from pure O2,fed through bubble diffusors;target O3concentration?1.4mg là1Lenz et al.(2004)

2.2m M O3generated from pure O2;fed at30ml minà1Kamiya et al.(2005) 88m M

20m M Anode:cobalt-coated PbO2;cathode:stainless steel;current density?0–50mA cmà2Korshin et al.(2006) 700m M Anode:platinum gauze;cathode:carbon felt;[Fe2+]/[BPA]0?1.42–7.14;[Cu2+]/[BPA]0?1.4–7.1Go zmen et al.(2003) 220m M US g frequency?500kHz,power?120W Kitajima et al.(2006) 110m M US bath,frequency?43–47kHz,power?500W;H2O2?7mg là1;FeSO4á7H2O?1.4,2.5mg là1;pH?4.0,

5.0,

6.5

Ioan et al.(2007)

a Medium pressure.

b Low pressure.

c Intensity.

d High pressure.

e Cyclodextrin.

f Microwave.

g Ultrasound.

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Table3

Identi?ed byproducts of BPA oxidation by various AOPs

AOP method Byproduct Reference

Indirect photolysis(with H2O2)Phenol Neamtu and Frimmel(2006a)

1,4-Dihydroxylbenzene

1,4-Benzoquinone

Acetate

Oxalate

Indirect photolysis(with Fenton’s reagent)4-Hydroxyacetophenone

methylbenzofuran

Katsumata et al.(2004)

p-Hydroquinone

Phenol

p-Quinone

Indirect photolysis(with ferric and

carboxylate ions)

BPA-o-catechol Zhou et al.(2004)

BPA-semiquinone

BPA-o-quinone

Indirect photolysis(with natural humic substances)Monohydroxylated BPA

p-hydroquinone

Zhan et al.(2006) 4-Isopropenylphenol

Glycerol

2-Hydroxy-propanoic acid

Photocatalysis with TiO2(immobilized

particles)

1,1-Ethenylidenebis-benzene Lee et al.(2004)

4-Isopropylphenol

4-tert-butylphenol

Phenol

Photocatalysis with TiO2(suspension)4-Isopropylphenol Watanabe et al.(2003) Photocatalysis with TiO2Plain Horikoshi et al.(2004)

(plain suspension,4-Hydroxyacetophenone

microwave irradiation,thermal)4-Hydroxyphenyl-2-propanol

Microwave

4-Hydroxyacetophenone

4-Hydroxybenzaldehyde

3-Hydroxy-1,3,5-hexadiene

Thermal

4-Hydroxyacetophenone

4-Hydroxyphenyl-2-propanol

4-Hydroxybenzaldehyde

Phenol

Hydroquinone

Photocatalysis with TiO2(suspension)3-(4-hydroxyphenyl)-3-methyl-2-oxobutanoic acid

4-vinylphenol

Ohko et al.(2001)

4-Hydroxyacetophenone

Photocatalysis with TiO2p-Hydroxyacetophenone Kaneco et al.(2004) (suspension,solar radiation)Methylbenzofuran

phenol

Photocatalysis with TiO2(suspension)p-Hydroxyacetophenone Fukahori et al.(2003)

p-Hydroxybenzaldehyde

p-Hydroxy-a-methylstyrene

hydroquinone

Ozonation p-tert-butylphenol

2-methyl-2,3-dihydrobenzofuran

hydroquinone

Gu ltekin et al.(unpublished data)

n-Butyl-acetate

Electrochemical oxidation(EC-generated Fenton’s reagent)m-Monohydroxylated BPA Go zmen et al.,2003. o-Monohydroxylated BPA

dihydroxylated BPA

Phenol

I.Gu¨ltekin,N.H.Ince/Journal of Environmental Management85(2007)816–832 824

and sonolysis.There is a wider scope of research on the degradability of NPEOs by AOP;however,the review presented in the following three sections is devoted to NP only for the fact that they are more toxic and more persistent than the ethoxylated form and they exhibit estrogenic activity unlike the ethoxylates.

3.2.1.Direct and indirect photolysis with UV

There is only one source of literature on direct photolysis of NP—the study of Neamtu and Frimmel(2006b),who investigated the decomposition of NP under a solar simulator equipped with a Xe lamp.They found that the rate of decay increased with pH elevations as a conse-quence of the larger photoreactivity of the deprotonated molecule;but it decreased with elevations in initial NP concentration.The effect of H2O2addition(indirect photolysis)was to accelerate the reaction,the intermediate byproducts of which were phenol and1,4-dihydroxylben-zene.Addition of bicarbonate and nitrate ions retarded and enhanced the rate of degradation upon excess consumption and production of OH radicals(by NO3àphotolysis),respectively.

Another study focusing on indirect photolysis was conducted by Chen et al.(2007)to evaluate the performance of a UV/H2O2process for the removal of endocrine activity in a mixed EDC solution containing NP.The ef?cacy of the process was assessed by monitoring EST using an in vitro yeast screen assay and an in vivo?sh assay.Kinetics of EST removal in both assays followed pseudo-?rst-order rate law regardless of the water matrix,but the rate was faster in deionized water than in river water,which the authors attributed to a lower steady state OH radical concentration in the latter due to excess consumption by dissolved organic constituents.They also reported that the rate of EST activity removal in the in vivo assay was faster than that in the in vitro,to be explained by differences in pharmacokinetic properties of the assays and to highlight the signi?cance of applying at least two tests for reliable results.

3.2.2.Photocatalysis

The majority of published research on advanced oxida-tion of NP is about photocatalytic methods,using plain or various modi?cations of TiO2and other catalysts.Pelizetti et al.(1989)showed that photocatalytic degradation and mineralization of4-NP is faster than that of phenol and p-n-propylphenol as a consequence of larger adsorption of NP on surfaces of TiO2pellets.They proposed that the reaction was initiated by d OH attack on the benzene ring resulting in the production of low molecular weight intermediates and CO2.Ike et al.(2002)compared heterogeneous and homogeneous degradation of NP by photocatalysis in TiO2suspension and ozonation,and reported90%and75–80%degradation in30and6min, respectively.

Kohtani et al.(2003)modi?ed the TiO2process by using a solar simulator and replacing the catalyst with BiVO4 (capable of splitting water to H2and O2)to study the degradation of4-n-NP.They found that the compound was adsorbed perpendicularly on BiVO4surface,while a mono-molecular layer covered over and the nonyl group faced to the surface.Thus,the long alkyl chain acted as the anchor on the hydrophobic BiVO4surface and thus zero-order kinetics predominated.Monitoring of the estrogenic activity showed that EST remained constant during the ?rst60min,decreased afterwards and disappeared com-pletely in140min.The major and minor oxidation byproducts were cis,cis-4-alkyl-6-oxo-2,6-hexadienoic acid and4-alkylcatechol and4-(1-alkenyl)phenol,respectively. In another study,the same research group investigated the effect of loading silver?ne particles on BiVO4and found that photocatalytic properties were strongly enhanced by the presence of silver oxides that partly covered the silver surface(Kohtani et al.,2005).

The study by Kurinobu et al.(2007)is about the photocatalytic decomposition of NP using?ne magnetic photocatalyst particles with a core–shell structure of three layers,Fe3O4,SiO2and TiO2.They found that in the absence of UV irradiation,NP was nearly all adsorbed on the magnetic particles during the?rst5min.Irradiation of the suspension by black light after120min resulted in complete NP removal.They also reported100% recovery of the particles(using a high gradient magnetic separation?lter)as an indicator of cost effectiveness of the process.

Table3(continued)

AOP method Byproduct Reference

Catechol

Hydroquinone

Benzoquinone

Resorcinol

4-Isopropenylphenol

4-Hydroxy mandelic acid

4-Hydroxy benzoic acid

Butendionic acid

4-Oxobutenoic acid

Acetic acid formic acid

Sonolysis2,3-Dihydro-2-methylbenzofuran Kitajima et al.(2006)

I.Gu¨ltekin,N.H.Ince/Journal of Environmental Management85(2007)816–832825

3.2.3.Dark advanced oxidation reactions

Studies on destruction of NP by dark advanced oxidation techniques such as ozonation are limited.Lenz et al.(2004)found that in a solution containing 1.4mg l à1ozone,the concentration of 4-NP (C 0?1–200m g l à1)could be reduced to less than 20ng l à1within 5min and the estrogenic activity to less than 95%of the original.A similar study by Kim et al.(2005)showed that despite the rapid breakdown of 4-NP by 1h ozonation,the concentration of dissolved organic carbon remained con-stant throughout.They also reported that total yield of aldehydes was directly related to the quantity of ozone in solution,while the predominance of acetaldehyde and formaldehyde over all other aldehydes was stable over all tested ozone concentrations.The kinetics of ozone-mediated oxidation of NP was studied by Deborde et al.(2005)with 4-n -NP over a pH ?2.5–10.5in the presence of t -butyl alcohol and reported that the rate was minimum at pH o 5and maximum at pH 410.The observation was justi?ed by estimation of the second-order rate constants,which showed that the rate was 4.5orders of magnitude larger at the alkaline level.A most recent study by Ning et al.(2007)also reports the kinetics of NP degradation by direct ozonation at pH ?2.The stoichiometric factor was estimated as 1.3(mol O 3:mol NP)and the apparent rate constant as 3.90?104M à1s à1,which is in good agreement with the value reported by Deborde et al.(2005)as 3.80?104M à1s à1.The reaction pathway proposed was initiated by the addition of one hydroxyl group at the ortho -position of the aromatic ring,leading to the formation of hydroxyl-alkylphenol.

Yim et al.(2003)investigated the degradation of nonylphenol in water by ultrasonic irradiation at 200kHz under argon,oxygen and air atmospheres and compared it

with the degradation of butyl-,pentyl-and octylphenols.They found that the rate of sonochemical oxidation was directly proportional to the length of the alkyl chain and fastest for all in argon-saturated solutions.Addition of a strong d OH scavenger (t -butanol)to solution was found to decelerate the reaction to signify the dominant role of OH radicals in the overall degradation process.However,addition of just suf?cient quantities of Fe(II)and Fe(III)under O 2atmosphere largely accelerated the degradation and mineralization of NP,which the authors attributed to the presence of dissolved O 2that acted as an extra source of OH radicals in the presence of ferric and ferrous catalysts.

3.2.

4.Overview of reaction conditions and oxidation byproducts

A comparative list of experimental conditions employed in studies reviewed so far for NP is given in Table 4.As stated for BPA,discrepancies in the observed data in different laboratories for a given process are due to differences in the applied conditions and the analyzed parameters.Note the large variation in the test concentra-tion,which is a major control parameter in the overall ef?ciency of an advanced oxidation system.Note also that we could not provide a table of byproducts list as we did for BPA,because byproducts as reported only in a couple of articles were pointed out in the text.

3.3.Destruction of phthalates by advanced oxidation processes

3.3.1.Direct and indirect photolysis with UV

Direct and indirect methods of photolysis have been more frequently investigated for destroying phthalates than

Table 4

Summary of reaction conditions for NP removal from water by AOP a Concentration

Reaction conditions

References

Direct and indirect photolysis with UV

25.5m M 1000W Xe short-arc lamp;H 2O 2?10,20,50mM;pH ?5.4,8.5;HCO 3à?725mg l à1

;

NO 3à?61mg l

à1

Neamtu and Frimmel (2006b)

0.18m M 15W LP Hg UV lamps (x4),(253.7nm);?uence ?0–2000mJ cm à2;H 2O 2?10mg l à1Chen et al.(2007)Photocatalysis 210m M

1500W Xe lamp (340nm);TiO 2?2g l à1

Pelizetti et al.(1989)0.9–14.5m M 4W LP Hg lamp (253.7nm);TiO 2?2g l à1;aeration ?4l min à1;O 3?17g m à3,?ow ?0.03l min à1

Ike et al.(2002)200m M 1000W Xe arc lamp (4400nm);I ?24mW cm à2;Catalyst:BiVO 4?4g l à1;pH ?13Kohtani et al.(2003)200m M Visible light (4400nm);I ?18mW cm à2;4g l à1Ag-loaded BiVO 4;pH ?13Kohtani et al.(2005)4.54m M

Black light (352nm),I ?2.6mW cm à2;TiO 2?5g l à1as magnetic particles

Kurinobu et al.(2007)Dark advanced oxidation reactions

4.54?10à3m M

0.908m M O 3generated from pure O 2;fed through bubble diffusors;target O 3concentration ?1.4mg l à1Lenz et al.(2004)2.8m M O 3stock solution produced from pure O 2;[O 3]/[NP]?0–10.4(M:M)Kim et al.(2005)4m M O 3stock solution produced from pure O 2-dissolved O 3?18mg l à1;pH ?2Ning et al.(2007)30m M US frequency ?200kHz,I ?6W cm à2;Fe(II)–Fe(III)?0–200m M

Yim et al.(2003)

a

Abbreviations ‘‘LP’’,‘‘I’’,‘‘US’’same as de?ned for Table 2.

I.Gu ¨ltekin,N.H.Ince /Journal of Environmental Management 85(2007)816–832

826

for NP.These methods consist of exposing the phthalate solution to short or medium wavelength UV radiation in the absence and presence of chemical reagents (e.g.ferric salts and ozone),identi?cation of the operating parameters,the reaction byproducts and/or determining the estrogenic activity.The following discussion covers a brief review of such literature.

Photodegradation by solar radiation is a major natural destruction pathway of non-biodegradable organic com-pounds in aquatic systems.A study devoted to the investigation of the degradability of di-n-propyl phthalate (DPP)by natural and arti?cial photolysis was conducted by Okamoto et al.(2006)and found surprisingly that both solar and arti?cial radiation(Hg arc)resulted in the formation of the same estrogenic byproduct.The impact of radiation wavelength on the reaction was further examined by using a Xe arc lamp with emissions at225,260,275,290 and310nm.It was found that estrogenic byproduct formation was a maximum with290nm irradiation and zero with225nm.They also reported that addition of H2O2 to enhance the reaction supported the formation of this undesired byproduct.

Lau et al.(2005)investigated direct photolysis of DBP under monochromatic UV irradiation at254nm over a pH range of3–11.They observed that the rate of photolysis slowed down after20–30min of irradiation by a factor related to the applied pH level.The retardation was attributed to the competition between intermediate pro-ducts and DBP for photons in the254nm band.The six major byproducts identi?ed were mono butyl phthalate (MBP),MBP-derived ketone(or aldehyde),MBP-derived alcohol,butyl benzoate,benzoic acid and phthalic acid,the latter being the dominant product at all test pH.The authors concluded that the applied method rendered both detoxi?cation and inhibition of endocrine activity. Another such study carried out by Oh et al.(2006)was focused on comparison of direct and indirect photolysis (with ozone)of DEP under254nm UV radiation with Hg low pressure lamps.The authors reported that while only 22%destruction was possible with direct photolysis, complete conversion of DEP was attained upon dosing 1.5mg là1minà1ozone to the photoreactor.The addition of t-butanol was found to lower the ef?ciency of the UV/O3 indirect system,showing the dominance of OH radicals in the degradation pathway.More than93%mineralization was reported for indirect method at an ozone dose of 4mg là1minà1.Monitoring of the EST activity showed that even after complete removal of the parent compound, the ef?uent exhibited weak EST activity due to the presence of residual byproducts.

The majority of indirect methods studied so far for phthalate destruction is about the Fenton reaction and variations of it.Bajt et al.(2001)investigated the photolysis of DBP in the presence of Fe(III)and reported85% destruction in90min irradiation by an arti?cial light source at365nm.The control parameter was the concen-tration of monomeric Fe(III)hydroxyl complex Fe(OH)2+due to its high photoactivity that resulted in the formation of additional hydroxyl radicals.The authors also reported that solar radiation was twice more effective than arti?cial light in the mineralization of the compound.The advan-tage of this process over conventional photo-Fenton processes is that there is no need for H2O2addition and the concentration of Fe(II)remains stable,acting as a continual source of OH radicals.A similar study with solar light-induced degradation of DEP in the presence of Fe(III)was conducted by Mailhot et al.(2002),who reported that the compound degraded only via OH radicals that were formed upon excitation of Fe(OH)2+as previously observed by Bajt et al.(2001),and the radiation source had no affect on the ef?ciency.Total mineralization after3-d irradiation was85%.The degradation of DEP was initiated by OH attack on the aromatic ring,while that of DBP involved a major and a minor attack at the alkyl chain and the aromatic ring,respectively(Mailhot et al., 2002;Bajt et al.,2001).Differences in reaction sites and reaction pathways are obviously due to differences in the length of the alkyl chain,which dictates the number of available reaction sites.

The study of Zhao et al.(2004)compares the ef?ciencies of dark and photo-Fenton processes for the degradation of DMP and reports that the compound was degraded by both of the processes,but with a2-fold faster rate by the light reaction.The optimum Fe+2and H2O2concentrations for the photo-Fenton process were1.67?10à4and5?10à4M, respectively,at which81%of DMP was oxidized in120min. The reported optimal reagent doses are in perfect agreement with the observation of Yang et al.(2005)for DEP,which was slightly less degradable(75.8%)under the same optimal conditions due to its larger alkyl chain.The most recent literature on the photo-Fenton process is the study of Chiou et al.(2006a)with DBP.It was found that at pH?3with a H2O2addition rate of4.74?10à5mol minà1là1and a Fe3+ concentration of 4.50?10à4mol là1,92%mineralization was possible in90min.

Finally,the study of Xu et al.(2007)is about the operating parameters of UV/H2O2process for the destruction of DEP using a low pressure Hg lamp emitting UV radiation at 254nm.The authors reported that at optimized conditions (UV intensity?133.9m W cmà2;H2O2?20mg là1),99%of the parent compound was removed in40min.The reaction rate was directly proportional to the concentrations of DEP and H2O2,and to the intensity of UV radiation and temperature.A direct linear relation between the reaction rate and the H2O2concentration is due to the fact that H2O2 was maintained below the critical level,above which it would compete with the target compound for OH radicals and retard its rate of oxidation(Ince,1999).

3.3.2.Photocatalysis with TiO2

The literature on photocatalytic destruction of phthalates describes variations in the form of TiO2catalyst,optimiza-tion of the operating parameters,and less frequently identi?cation of the reaction byproducts.The earliest study

I.Gu¨ltekin,N.H.Ince/Journal of Environmental Management85(2007)816–832827

is about the degradability and oxidation byproducts of dialkyl phthalates in TiO2suspension(Hustert and Moza, 1988).The authors found that mono and di-hydroxylated byproducts were formed upon oxidation of phthalates containing shorter alkyl chains(dimethyl and diethyl), whereas hydroxylated byproduct formation was insigni?cant in the oxidation of phthalates with longer alkyl chains (dibutyl and di-ethyl-hexyl).In a recent study by Kaneco et al.(2006),the same byproducts were identi?ed for the destruction of DBP in TiO2suspension and found that optimum values of the catalyst concentration,temperature and pH were50m g mlà1,201C,and6.0,respectively. Muneer et al.(2001)have also focused on setting of the operation parameters such as pH,concentration and type or quantity of TiO2particles for the degradation of DEP. The ef?ciency de?ned as the ratio of the degradation rate to the incident light intensity was found to increase with pH increase and reached a maximum at pH6,which is close to the zero point charge of TiO2.The authors concluded that DEP conversion ef?ciency could be enhanced by raising the concentration of TiO2and/or that of DEP.Although the addition of electron acceptors such as hydrogen peroxide,bromate and persulfate did not render a signi?cant enhancement in the rate of DEP conversion,the degree of mineralization was remarkably increased in the presence of bromate.A more recent study describing the use of TiO2particles immobilized on silicate glass beads,as reported by Chiou et al.(2006b)for DBP showed that75%conversion and70%mineralization was possible in80min regardless of the pH level.Hence,unlike most studies reporting pH as a critical parameter,the in?uence of pH within a range 4.5–9.0was found insigni?cant in this study,as attributed to the scavenging of excess OH radicals(generated in alkaline pH)by carbonate species.

The study of Ooka et al.(2003)compared the ef?ciency of untreated and hydrothermally treated TiO2pillared clays on the degradation of DBP,DEP,DMP and BPA,reporting that the process was enhanced with increasing hydrophobi-city of the phtalate(DBP4BPA4DEP4DMP)and with hydrothermal treatment of the clays.In a subsequent study, the authors further compared the effectiveness of four different hydrophobic TiO2pillared clays on the degrada-tion of DBP and DMP,and found that the degradation of both compounds was enhanced with increasing surface hydrophobicity of the clays,by which the adsorptive capacity also increased(Ooka et al.,2004).

3.3.3.Dark advanced oxidation reactions

The literature on dark AOPs of phthalate destruction is limited to ozonation and ultrasound processes.A study by Li et al.(2005)showed that the rate of ozone-based destruction of DBP is directly proportional to the applied ozone dose and is fastest during the?rst15min of reaction. The authors found that mineralization was much slower than oxidation of the parent compound,and attributed it to the formation of organic acids with low reactivity with ozone.Addition of UV light and/or UV/TiO2increased both the rate of degradation and the fraction of mineraliza-tion upon photo-decomposition of O3and/or photoactiva-tion of TiO2,respectively,to generate active oxygen species. There is also research with catalytic ozonation processes to enhance the biodegradability and/or the mineralization of phthalates.One such study compares the ef?ciency of a bio-treatment process applied to the ef?uent of a plain ozonation and a granular-activated carbon-catalyzed ozonation process for the treatment of DMP,DEP,DBP and DEHP(Li et al.,2006).The authors found that pre-ozonation with activated carbon improved ozone utiliza-tion and the biodegradability of the ef?uent,resulting in nearly100%removal of DMP and DEP(and more than 93%of the others)in the ef?uent of the bio-process.More recently,Yunrui et al.(2007)have compared the ef?cien-cies of Al2O3and Ru/Al2O3–catalyzed ozonation processes for the destruction of DMP.They found the catalysts were ineffective in the oxidation of DMP by ozone,but very effective in the mineralization of the ozonated byproducts, as the CO2formation increased by3-fold and1.5-fold in the presence of Ru/Al2O3and Al2O3,respectively.Con-tinuous monitoring of the total organic carbon and component leaching for42h showed that catalytic activity of Ru/Al2O3remained constant with time,while that of Al2O3decreased upon leaching.

There are two published articles in the literature about ultrasonic destruction of phthalates.The earlier one is by Yim et al.(2002),who investigated the degradation of phthalic acid esters(DMP,DEP and DBP)at200kHz. They reported that the rate of all was pseudo-?rst order with half-lives of17.9,11.8and6.9min,respectively,and the reaction site was either the gas–liquid interface (pH4–11)or the cavitation bubble(pH411).More recently,Psillakis et al.(2004)studied the decomposition of six phthalate esters at80kHz.They showed that higher molecular mass phthalates such as DBP,BBP,DEHP and di-octyl phthalate(DOP)were readily degradable by ultrasound for complete removal in30–60min contact, while less hydrophobic DMP and DEP were much more resistant,requiring prolonged sonication for complete destruction.Addition of NaCl was found to enhance the rate of degradation in more polar DMP,DEP,DBP and BBP upon decreased solubility,which promotes diffusion of the compounds from the bulk solution to the bubble–water interface,where they meet a much denser cloud of hydroxyl radicals.On the other hand,salt addition was reported to decelerate the degradation of non-polar esters such as DEHP and DOP to be explained by reduced number of cavity bubbles upon alteration of the vapor pressure.

3.3.

4.Overview of reaction conditions and oxidation byproducts

The experimental conditions referring to the articles discussed in Sections 3.3.1–3.3.3are summarized in Table5.Similar to the conditions of BPA and NP,the largest variation is in the concentration of the test

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compound.There are also vast variations in the applied

light intensity in UV-based reactions and in the concentra-tion of the TiO 2catalyst.However,the concentration of iron and H 2O 2in Fenton processes lies within same orders of magnitude.Byproduct analysis was rarely carried out for phthalate destruction by AOP,and the only informa-tion is the formation of monohydroxylated and/or dihydroxylated derivates (Hustert and Moza,1988;Bajt et al.,2001;Mailhot et al.,2002;Kaneco et al.,2006).4.Conclusions

The increasing consumption of EDCs worldwide has raised signi?cant public concern due to their initiation of hormone-like activities even in trace concentrations in surface waters.Research has shown that AOPs,which

generate very active oxidative species such as the hydroxyl radical,are promising tools for the destruction of such compounds in water,particularly bisphenol-A,nonylphe-nol and phthalates.The most commonly investigated method of advanced oxidation for destroying EDCs in water is photocatalysis with titanium dioxide,while there is suf?cient literature on ozonation,direct and indirect photolysis,electrochemical oxidation and ultrasonic irra-diation,as well.However,more research is required particularly with hybrid processes to make advantage of synergistic effects and to investigate the impacts of interfering agents and chemical structures before proposing advanced oxidation as a viable solution to the treatment of these compounds in water.

A large majority of the reviewed articles involved studying the behavior of a single chemical tested in much

Table 5

Summary of reaction conditions for phthalate removal from water by AOP a Compound/concentration

Reaction conditions

Reference

Direct and indirect photolysis with UV DPP/250m M Solar,arti?cal UV irradiation (Hg arc-254nm,0.94mW cm à2or Xe arc:225,260,

275,290,310nm,0.07,0.16,0.28,0.34,0.48mW cm à2;H 2O 2?15m M

Okamoto et al.(2006)DBP/2–10m M 35W–254nm phosphorcoated LP Hg lamps (x8),I ?1.5?10à6Einstein l à1s à1;

pH ?3–11.

Lau et al.(2005)DEP/100m M 3W LP Hg arc lamp (254nm);I ?0.4W l à1;O 3dose ?1.5,4mg l à1min à1,fed at

0.5l min à1;pH ?7

Oh et al.(2006)DBP/30m M 125W HP Hg lamp (365nm);Fe(III)?3?10à4M;pH ?3.1Bajt et al.(2001)DEP/260m M 125W HP Hg lamp (365nm)or natural sunlight;Fe(III)?3.05?10à4M Mailhot et al.(2002)DMP/51.5m M 150W HP Hg lamp;Fe(II)?0–3.3?10à4M;H 2O 2?0–8.3?10à4M;pH ?2–6Zhao et al.(2004)DEP/45m M 160W HP Hg lamp;Fe(II)?0–3.3?10à4M;H 2O 2?0–10?10à4M;pH ?1–6Yang et al.(2005)DBP/18m M 8W UV (312nm)lamps (x2);I ?120m W cm à2;pH ?2–4,Fe(III)

?9–54?10à5M,H 2O 2?4–55?10à6M min à1;T ?251C

Chiou et al.(2006a)DEP/4.5m M 30W (254nm)UV lamps(x10);I ?21–134m W cm à2;H 2O 2?2.5–30mg l à1;

T ?15–31%oC

Xu et al.(2007)

Photocatalysis with TiO 2DBP/18m M 990W Xe lamp (o 300nm);I ?1.8mW cm à2;TiO 2?0–0.3g l à1;pH ?1–10;T ?10–401C

Kaneco et al.(2006)DEP/100–1000m M 500W HP Hg lamp (320nm);I ?110m mol photon l à1min à1;TiO 2?0.5–5g l à1;pH ?3–9

Muneer et al.(2001)DBP/9–45m M

8W UV lamp (365nm);I ?1.67mW cm à2;150g TiO 2/glass;pH ?4.5–9.0;sample circulation ?10ml min à1

Chiou et al.(2006b)DBP/144m M,DEP/180m M,DMP/206m M Black light lamp (365nm);I ?0.80mW cm à2;TiO 2?1.2g l à1Ooka et al.(2003)DBP/144m M

Black light lamp (365nm);I ?0.80mW cm à2;TiO 2?1.2g l à1

Ooka et al.(2004)

DMP/206m M

Dark advanced oxidation reactions with ozone and ultrasonic processes DBP/54m M Ozone feed ?12.5,25.0,50.0mg h à1;pH ?6.4;15W LP UV lamp (254nm);

I ?40mW cm à2;carbon-modi?ed TiO 2thin ?lms

Li et al.(2005)DMP/593m M,DEP/754m M,DBP/106.8m M,DEHP/23m M O 3generated from pure O 2;100g GAC in ozone reactor;EBCT-BAC b reactor ?15min Li et al.(2006)DMP/31m M Ozone feed ?118mg O 3‘h à1;gas ?ow ?400ml min à1;Ru/Al 2O 3

catalyst ?10g l à1;pH ?6.6,T ?151C

Yunrui et al.(2007)DMP,DEP,DBP/100m M US frequency ?200kHz;US power I ?6W cm à2;pH ?4–12Yim et al.(2002)DMP/0.206m M,DEP/0.180m M,DBP/0.144m M,BBP/0.128m M,DEHP,DOP/0.102m M

US frequency ?80kHz;US power ?75W,150W Psillakis et al.(2004)

a Abbreviations ‘‘LP’’,‘‘HP’’,‘‘I’’,‘‘US’’same as de?ned for Table 2.b

Empty bed contact time in biological activated carbon reactor.

I.Gu ¨ltekin,N.H.Ince /Journal of Environmental Management 85(2007)816–832

829

higher concentrations than detected in the water environ-ment,whereas endocrine disruptors exist in mixtures of multi components rather than as a single component.In present,the dif?culty of working with concentrations in micro and nano gram là1levels to simulate the environ-ment is due to the limitations in available analytical instruments,but the problem will be solved in the future with promising developments in instrumentation and detection limits.As a?nal remark,it should be noted that recent developments in bioassay techniques used for screening the ef?uent endocrine activity provide excellent support to the evaluation of the overall performance of the treatment processes.

Acknowledgments

The authors thank Bogazici University Research Fund for funding the Project06S101.

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