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Degradation of Aqueous Pharmaceuticals by Ozonationand Advanced Oxidation Processes: A ReviewKeisuke Ikehata a; Naeimeh Jodeiri Naghashkar a; Mohamed Gamal El-Din aa Department of Civil and Environmental Engineering, CNRL Natural ResourcesEngineering Facility, University of Alberta, Edmonton, Alberta, Canada

Online Publication Date: 01 December 2006To cite this Article: Ikehata, Keisuke, Naghashkar, Naeimeh Jodeiri and El-Din,Mohamed Gamal (2006) 'Degradation of Aqueous Pharmaceuticals by Ozonationand Advanced Oxidation Processes: A Review', Ozone: Science & Engineering,

28:6, 353 - 414To link to this article: DOI: 10.1080/01919510600985937URL: http://dx.doi.org/10.1080/01919510600985937

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Ozone: Science and Engineering, 28: 353–414

Copyright # 2006 International Ozone Association

ISSN: 0191-9512 print / 1547–6545 online

DOI: 10.1080/01919510600985937

Degradation of Aqueous Pharmaceuticals by Ozonationand Advanced Oxidation Processes: A Review

Keisuke Ikehata, Naeimeh Jodeiri Naghashkar, and Mohamed Gamal El-Din

Department of Civil and Environmental Engineering, CNRL Natural Resources Engineering Facility, University of Alberta,Edmonton, Alberta, Canada

A vast number of pharmaceuticals have been detected insurface water and drinking water around the world, whichindicates their ineffective removal from water and waste-water using conventional treatment technologies. Concernshave been raised over the potential adverse effects of phar-maceuticals on public health and aquatic environment.Among the different treatment options, ozonation andadvanced oxidation processes are likely promising for effi-cient degradation of pharmaceuticals in water and waste-water. Recent progress of advanced oxidation of aqueouspharmaceuticals is reviewed in this paper. The pharmaceu-ticals and non-therapeutic medical agent of interest includeantibiotics, anticonvulsants, antipyretics, beta-blockers,cytostatic drugs, H2 antagonists, estrogenic hormone andcontraceptives, blood lipid regulators, and X-ray contrastmedia.

Keywords Ozone, Advanced Oxidation Processes, Antibiotic,Antipyretic, Anticonvulsants, Blood Lipid Regula-tor, Beta-blocker, Cytostatic Drug, X-Ray ContrastMedia, Estrogens

INTRODUCTION

A large amount of numerous prescription and non-prescription drugs of different classes are consumedannually throughout the world. These pharmaceuticalcompounds include antipyretics, analgesics, blood lipidregulators, antibiotics, antidepressants, chemotherapyagents, and contraceptive drugs. After the administration,these compounds are partially metabolized and excreted

in the urine and feces, and subsequently enter into sewagetreatment plants where these compounds are treated,along with other organic and inorganic constituents inwastewater. However, it has been shown that some ofthese pharmaceutical compounds are not completelyremoved in sewage treatment plants (Halling-Sørensenet al., 1998; Ternes, 1998; Ternes et al., 1999; Heberer,2002; Boyd et al., 2003). As a result, they have been foundin some sewage treatment plant effluents as well as insurface water and groundwater in many countries(Ternes, 1998; Sacher et al., 2001; Soulet et al., 2002;Miao et al., 2004). In addition, non-therapeutic medicalagents such as X-ray contrast agents (also known asradiocontrasts) and some veterinary drugs have beenalso found in the aquatic environment (Ternes andHirsch, 2000; Drewes et al., 2001; Kummerer, 2001).Although some of these pharmaceuticals and their meta-bolites can be partially removed through sorption andbiotic or abiotic degradation in the environment, theycan eventually reach drinking water sources. Severalrecent studies have revealed that conventional watertreatment processes cannot remove some prescriptionand non-prescription drugs completely from sourcewaters (Ternes et al., 2002; Stackelberg et al., 2004;Jones et al., 2005a).

The frequent occurrence of pharmaceuticals in theaquatic environment as well as in finished drinkingwater has raised a concern about their potential impacton environmental and public health. Some of the adverseeffects caused by pharmaceutical pollution include aqua-tic toxicity, resistant development in pathogenic bacteria,genotoxicity, and endocrine disruption (Arcand-Hoyet al., 1998; Halling-Sørensen et al., 1998; Sumpter,1998; Kummerer, 2004). The presence of trace pharma-ceutical and other xenobiotic compounds in finisheddrinking water is another public health concern, sincelittle is known about potential chronic health effects

Received 03/28/2006; Accepted 06/05/2006Address correspondence to Mohamed Gamal El-Din, Dept. of

Civil and Environmental Engineering, 3-093 Markin/CNRLNatural Resources Engineering Facility, University of Alberta,Edmonton, Alberta T6G 2W2, Canada. E-mail: mgamalel-din@ualberta.ca

Ozonation and Advanced Oxidation of Pharmaceuticals December 2006 353

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associated with long term ingestion of mixtures of thesecompounds through drinking water (Kummerer, 2001;Stackelberg et al., 2004). Thus, it is an emerging issue inenvironmental science and engineering to achieve effec-tive removal of pharmaceutical compounds, along withother priority pollutants, from the sources such as hospi-tal and domestic wastewaters before their discharge.

As conventional water and wastewater treatment pro-cesses are unable to act as a reliable barrier toward someof recalcitrant pharmaceuticals, it is necessary to intro-duce additional advanced treatment technologies in theareas where a persistent pollution problem has beenrecognized or is anticipated. Various advanced treatmenttechnologies have been evaluated for this purpose inrecent years, including chemical oxidation using ozoneand ozone/hydrogen peroxide (Zwiener and Frimmel,2000; Ternes et al., 2002), membrane filtration such asnanofiltration and reverse osmosis (Hartig et al., 2001;Heberer et al., 2002; Nghiem et al., 2004; Nghiem et al.,2005), and activated carbon adsorption (Hartig et al.,2001; Ternes et al., 2002).

The latter two processes are highly energy and materialintensive (Larsen et al., 2004) and only suitable for thetreatment of relatively clean surface water and ground-water with less background contamination such as naturalorganic matter (NOM). These physical treatment processesalso require the disposal of wastes such as membraneretentate and spent activated carbon generated duringthe treatment. In addition, activated carbon adsorptionhas a limited ability to remove polar organic compoundsdue to its removal mechanism (i.e., hydrophobic inter-actions), especially in the presence of competitive NOM(Snyder et al., 2003), and many pharmaceutical com-pounds and metabolites are indeed polar substances. Onthe other hand, chemical oxidation such as ozonation andadvanced oxidation processes (AOPs) may be a moreappropriate treatment option for pharmaceutical com-pounds in wastewater as well as in surface water andgroundwater.

In this review, recent literature concerning ozonationand advanced oxidation treatment of a number of phar-maceutical compounds in aqueous medium are reviewedand discussed in terms of the degree of reaction (i.e.,degradation), reaction kinetics, identity and characteris-tics of oxidation by-products, and possible degradationpathways. Nine major classes of pharmaceuticals listed inTable 1 are covered, including antibiotics, anticonvul-sants and antidepressants, antipyretics and non-steroidalanti-inflammatory drugs, beta-blockers, cytostatic drugs,histamin H2-receptor antagonists, hormones and oralcontraceptives, lipid regulators, and X-ray contrastmedia. A brief review of pharmaceuticals and its metabo-lism, their occurrence and fate in the environment, andtheir potential impact on environmental and public healthis provided in the following section as well. Please alsorefer to the comprehensive surveys published elsewhere

for details of these issues (Halling-Sørensen et al., 1998;Kummerer, 2001; Heberer, 2002; Snyder et al., 2003;Debska et al., 2004).

BACKGROUND

Pharmaceuticals and Their Metabolites

Pharmaceuticals, or medical drugs, are a group ofchemical substances that have medicinal properties.These substances include both inorganic and organiccompounds, although most of the modern pharmaceuti-cals are small organic compounds with a molecularweight below 500 Daltons (Lipinski et al., 1997). Thesechemicals are moderately water soluble as well as lipo-philic to be bioavailable and biologically active.Pharmaceuticals are either administered topically (e.g.,inhalation and skin application), internally (e.g., oraladministration), or parenterally (e.g., injections and infu-sions) in households or healthcare facilities such as hos-pitals and clinics. After the administration, drugmolecules are absorbed, distributed, metabolized, andfinally excreted from the body. To be used safely mostof the modern pharmaceuticals are designed in a way toundergo metabolism in organs such as liver and kidneyafter they have achieved desired pharmacological effects.Metabolisms detoxify excess drug molecules, as well asother toxic xenobiotics, via a series of enzymatic biotrans-formations and convert them to be more polar andhydrophilic (Silverman, 1992). The drug metabolismbegins with various biochemical reactions includinghydroxylation, epoxidation, reduction, and hydrolysis,in which functional groups are introduced or unmasked(phase I transformation). Afterwards, highly polar endo-genous molecules such as glucuronic acid, sulfate, andamino acids are attached to drugs or metabolites ofphase I transformation to generate conjugates (phase IItransformation) that are water-soluble and can be readilyexcreted in the urine or bile. There are certain drugs, non-therapeutic medical agents, and xenobiotics that are non-metabolizable because they are poor substrates for meta-bolizing enzymes such as cytochrome P-450. These com-pounds may be eliminated slowly from the body withoutbiotransformation.

Occurrence and Fate of Pharmaceuticals in theEnvironment

Until recently pharmaceutical compounds in the envir-onment have drawn very little attention. Although theirpresence in sewage treatment plant effluents was reported(Richardson and Bowron, 1985), it had been anticipatedthat these compounds were easily biodegradable inenvironment as most of them could be metabolized andtransformed to some extent in humans, as discussedbefore (Kummerer, 2001; Debska et al., 2004). However,a large number of recent studies have demonstratedpersistence of these pharmaceuticals in the aquatic

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environment. The occurrence of several pharmaceuticalcompounds have been reported in sewage treatment planteffluents as well as in surface waters in Germany (Ternes,1998; Hirsch et al., 1999; Putschew et al., 2000), theNetherlands (Belfroid et al., 1999), Switzerland (Souletet al., 2002), Canada (Ternes et al., 1999; Miao et al.,2004), Brazil (Ternes et al., 1999), Italy (Castiglioni et al.,2004), Spain (Rodrıguez et al., 2003), and the UnitedStates (Drewes et al., 2001; Kolpin et al., 2002). Thedetected compounds included antibiotics, anticonvul-sants, painkillers, cytostatic drugs, hormones, lipid regu-lators, beta-blockers, antihistamines, and X-ray contrastmedia. The concentrations of these pharmaceuticals werein the range of ng/L to mg/L in sewage treatment planteffluents and surface water. In addition, a number ofpolar pharmaceutical compounds and metabolites, suchas diclofenac, carbamazepine, sulfamethoxazole, andamidotrizoic acid, have been detected in groundwatersamples at concentrations up to 1 mg/L (Sacher et al.,2001; Cahill et al., 2004; Clara et al., 2004).

There are several possible sources and routes for theoccurrence of pharmaceutical compounds in the aquaticenvironment (Figure 1). For human pharmaceuticals,non-prescription drugs and some prescription drugs areconsumed in households, and other prescription drugs areconsumed in healthcare facilities such as hospitals and

clinics. These drugs are partially metabolized andexcreted in the urine and feces and go into a wastewatercollection system (Heberer, 2002; Jones et al., 2005b).Some unused, surplus, or expired drugs may be disposedinto toilets, although this kind of practice is not recom-mended nowadays. Wastewater from the hospitals maybe treated separately or combined with municipal waste-water and then treated at sewage treatment plants.

Some of the pharmaceuticals and (human) metabolitesin wastewater are degraded completely or partially, givingrise to a mixture of parent compounds and a variety ofmicrobial metabolites (Ternes, 1998; Drewes et al., 2001;Miao et al., 2002; Soulet et al., 2002; Jones et al., 2005b).Some pharmaceuticals such as ibuprofen and bezafibrateare relatively biodegradable, while others such as carba-mazepine and diazepam are practically non-biodegradable(Larsen et al., 2004). It is also known that some drugconjugates such as glucoronides can be cleaved by micro-bial degradation resulting in a release of parent com-pounds (Heberer, 2002; Jones et al., 2005b). Effluentfrom sewage treatment plants may be released to surfacewater or be subjected to groundwater recharges, so thatthe mixture of compounds enters the aquatic environment.In some cases, biologically treated municipal wastewatermay be treated further to produce various reclaimedwaters for different purposes including portable re-use.

TABLE 1. Pharmaceuticals Reviewed

Class Sub-class Pharmaceuticals covered

b-Lactam Amoxicillin, cefradine, ceftriaxone, penicillin, penicillin G,penicillin VK, sultamicillin, MMTD, MMTD-Me

Macrolide Azithromycin, clarithromycin, erythromycin,roxithromycin, lincomycin

Antibiotic Quinolone Ofloxacin, enrofloxacinSulfonamide Sulfachlorpyridazine, sulfadiazine, sulfadimethoxine,

sulfamerazine, sulfamethazine, sulfamethizole, sulfamethoxazole,sulfamoxole, sulfasulfapyridine, sulfathiazole, sulfisoxazole

Other Carbadox, spectinomycin, tetracycline, trimethoprimAnticonvulsantand antidepressant

Buspirone, carbamazepine, diazepam, primidone

Antipyretic and NSAID* Diclofenac, ibuprofen, indomethacin, naproxen, paracetamol,salicylic acid

b-blocker Atenolol, celiprolol, metoprolol, propranolol, sotalolAlkylating agent Cyclophophamide, ifosfamide, melphalan

Cytostatic drug Anthracycline Aclarubicin, daunorubicin, doxorubicin, epirubicin, idarubicin,pirarubicin

Anti-metabolite Azathioprine, cytarabine, 5-fluorouracil, methotrexateHistamine H2-receptorantagonist

Cimetidine, ranitidine

Hormone and oralcontraceptive

17b-Estradiol, estrone, 17a-ethinylestradiol, diethylstilbestrol

Lipid regulator Bezafibrate, clofibric acid, fenofibric acid, gemfibrozilX-ray contrast medium Diatrizoate, iomeprol, iopamidol, iopentol, iopromide

Note: * non-steroidal anti-inflammatory drug.

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Sorption can also occur during the sewage treatmentprocesses and some of the pollutants can be transferred tosewage sludge (Larsen et al., 2004). Sewage sludge may besubjected to anaerobic or aerobic digestion, conditioningand dewatering, and is subsequently landfilled, inciner-ated, or applied to lands as a biosolid. Pharmaceuticalcompounds in the sludge can be degraded further duringthe digestion, although some of them may remain intact.These compounds can be seeped into groundwater aqui-fers or be flushed by surface water runoff after the landapplication and can cause additional contamination pro-blems in the aquatic environment. Therefore, the sorptionof pharmaceuticals during sewage treatment processescannot be counted as a removal process unless the sludgewas incinerated (Larsen et al., 2004).

In addition to the drugs and medical agents for humanuse, a large amount of pharmaceutical compounds, espe-cially antibiotics, are used for various agricultural pur-poses, such as veterinary therapeutics, growth promotionof livestock, and feed additives in fish farms (Halling-Sørensen et al., 1998). As shown in Figure 1, these phar-maceuticals can enter into the aquatic environmentdirectly (feed additives for fish) or indirectly through live-stock manure (growth promoters and therapeutics). Land

application of livestock manure is often practiced andmay cause contamination problems in surface water andgroundwater similar to the case of municipal sewagesludge disposals described above. Additional sources ofpharmaceuticals in the aquatic environment include was-tewater from pharmaceutical production/formulationplants and waste disposal of unused drugs as a solidwaste.

ENVIRONMENTAL AND PUBLIC HEALTH IMPACTSOF PHARMACEUTICALS IN THE ENVIRONMENT

After publication of the earlier occurrence data, therisks associated with pharmaceutical contamination ofthe aquatic environment have become a major issue ofconcern for environmental scientists and engineers, aswell as among the public. Drugs are the chemicals thatare designed to give a certain therapeutic (=biological)effect; therefore, certain environmental and public healthrisks can be anticipated from the exposure to the envir-onmental pharmaceuticals. Besides, there are a fewclasses of pharmaceuticals that pose unambiguousimpacts on the aquatic organisms, including microorgan-isms, phytoplankton, plants, crustaceans, fish, and

FIGURE 1. Routes of pharmaceutical contamination of the aquatic environment.

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insects, as well as on soil microorganisms and possiblyhumans (Halling-Sørensen et al., 1998; Sumpter, 1998;Kummerer, 2001, 2004). These pharmaceutical classesinclude:

� Cytostatic agents, immunosuppressive drugs,and some genotoxic antibiotics because oftheir evident cytotoxic, carcinogenic, muta-genic, and/or embryotoxic properties;

� Human and veterinary antibiotics because oftheir pronounced microbial toxicity and thedevelopment of antibiotics resistance in envir-onmental bacteria including human pathogens;

� Natural and synthetic hormones because oftheir high efficiency, low effect thresholds andpotential for endocrine disruption;

� Halogenated compounds such as iodinated X-ray contrast media because of their resistancetoward biodegradation and their mobility andpersistence in the environment and the foodweb;

� Heavy-metal containing drugs and non-thera-peutic medical agents because of the toxicity ofthe metals in certain oxidation states.

In addition, the presence of other types of pharmaceu-ticals, such as analgesics and anticonvulsants, in drinkingwater is a potential public health issue. Although theconcentrations found in finished water is generally verylow, it is apparent that drinking water consumption is themajor route of human exposure to the environmentalpharmaceuticals (Figure 1). Since the long-term healtheffects are still largely unknown for the exposure to thetrace pharmaceuticals and their metabolites, especially asa mixture of biologically active compounds, the existenceof these compounds in drinking water should be avoidedon the basis of precautionary principle (Snyder et al.,2003; Jones et al., 2005a). Similarly, long-term exposureof aquatic organisms to trace pharmaceuticals in surfacewater may have some as-yet-known ecological impacts.

Control Measures for Pharmaceutical Contamina-tions in The Aquatic Environment

As shown in Figure 1, there are several routes ofpharmaceutical contamination of the environment. Thisimplies that a number of opportunities exist to control thepollution problems. However, as suggested by severalgroups of researchers (Larsen et al., 2004; Jones et al.,2005b), the conventional end-of-pipe approach can bevery costly and may not be a feasible option, and it ishighly desired to treat and remove those potentiallyhazardous pharmaceutical compounds in wastewaterand solid waste properly and as close as their primarysources. In the view of the aquatic environment, waste-water treatment is considered as the key step, at least tokeep out human pharmaceuticals. Because current sys-tems cannot remove some of pharmaceuticals effectively,

some improvement and modification will be necessary toaccount for this problem (Jones et al., 2005b). For exam-ple, increasing solids retention time in biological treat-ment processes will facilitate the development ofpopulation of slower growing bacteria and may enablethem to be acclimated to the recalcitrant compounds.Application of the advanced treatment technologies isanother option. Those advanced technologies includemembrane filtration (reverse osmosis and nanofiltration),activated carbon adsorption, ozonation, and AOPs(Petrovic et al., 2003). Although they are effective, nearlyall of these advanced technologies are energy and/ormaterial intensive to be applied to wastewater treatment,especially membrane processes and activated carbonadsorption. Introduction of ozonation or AOPs beforeor after biological treatment process may be feasiblebecause the chemical/photochemical oxidation rendersrecalcitrant xenobiotics more biodegradable and lesstoxic and improves their degradation in the followingtreatment process or in the environment (Alvares et al.,2001).

Potable water treatment is the last line of defensetoward the contamination of water for human consump-tion. Therefore, it is still necessary to ensure that finisheddrinking water is free from any potentially hazardoussubstances, including trace pharmaceuticals, to humans.This is particularly important in such areas where: (1)conventional drinking water sources are scarce and recla-mation of wastewater is required to supplement the watersources; (2) municipal sewage treatment facilities do notprovide effective removal of potentially toxic pharmaceu-ticals; and (3) major pollution sources such as sewagetreatment plants, farmlands, and pharmaceuticals manu-facturing/formulation plants, are located nearby. Theindividual or a combination of advanced treatment tech-nologies mentioned above can be applied to drinkingwater treatment as well.

In addition, source separation is also an importantmanagement strategy to alleviate the problem of pharma-ceutical pollution of the environment. Some of the phar-maceuticals are not usually consumed in households, butmainly in healthcare facilities. Those compounds includecytostatic agents, immunosuppressive drugs, some anti-biotics, and contrast media. On the other hand, someantibiotics, hormones, and many other prescription andnon-prescription drugs are widely consumed in the house-holds. Separated treatment of highly contaminated andpotentially more toxic hospital wastewater is probablydesired. Separating human urine from the rest of domes-tic wastewater is also considered as an attractive optionfor improving the water pollution control with respect tonutrients and micropollutants, including pharmaceuticals(Larsen et al., 2004). Most of xenobiotic compounds,including pharmaceuticals, are excreted by kidney aspolar, water-soluble metabolites. Treatment of pharma-ceuticals and their metabolites in the urine prior to

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dilution can be cost effective owing to the simple watermatrix as compared with combined wastewater (i.e., theabsence of interferences such as suspended solids andother dissolved organics). Chemical oxidation such asozonation and AOPs can be a viable treatment optionfor the separated urine.

Ozonation and Advanced Oxidation Treatment ofWater and Wastewater

Ozonation and AOPs have recently emerged as animportant class of technologies for the oxidation anddestruction of a wide range of organic pollutants inwater and wastewater (Legrini et al., 1993; Alvareset al., 2001; Zhou and Smith, 2001; Oppenlander, 2003).The AOPs are characterized by a variety of radical reac-tions that involve combinations of chemical agents (e.g.,ozone (O3), hydrogen peroxide (H2O2), transition metals,and metal oxides) and auxiliary energy sources (e.g.,ultraviolet-visible (UV-Vis) radiation, electronic current,g-radiation, and ultrasound). Examples of AOPs includeO3/H2O2, O3/UV, O3/H2O2/UV, H2O2/UV, Fenton(Fe2+/H2O2), photo- and electro-Fenton, chelatingagent assisted Fenton/photo-Fenton, heterogeneousphotooxidation using titanium dioxide (TiO2/hn), g-radi-olysis, and sonolysis (Oppenlander, 2003).

Hydroxyl radicals (�OH) are the primary oxidant inAOPs while other radical and active oxygen species suchas superoxide radical anions (O2

��); hydroperoxyl radi-cals (HO2

�); triplet oxygen (3O2), and organic peroxylradicals (R-O-O�) are also involved (Oppenlander,2003). Molecular ozone and hydroxyl radicals areinvolved in ozone treatment of water and wastewater.Ozonation at high pH (>8) is also considered as anAOP because of the enhanced generation of hydroxylradicals under such conditions (Beltran, 2003). The oxi-dation potentials of molecular ozone and hydroxyl radi-cals are 2.07 and 2.80 V, respectively, indicating that theyare very strong oxidants (Legrini et al., 1993). Whereasmolecular ozone reactions are selective to the organicmolecules having nucleophilic moieties such as carbon-carbon double bonds, aromatic rings, and the functionalgroups bearing sulfur, phosphorus, nitrogen and oxygenatoms, hydroxyl radicals reactions are non-selectivetoward various organic and inorganic compoundsthrough hydrogen abstraction, radical-radical reactions,electrophillic addition, and electron transfer reactions,and eventually lead to complete mineralization of organiccompounds (Legrini et al., 1993; Oppenlander, 2003).

Ozonation and AOPs are particularly appropriate formunicipal and industrial wastewaters containingbio-refractory and/or toxic organic pollutants such aspesticides (Reynolds et al., 1989; Ikehata and GamalEl-Din, 2005; Ikehata and Gamal El-Din, 2006), surfac-tants (Ikehata and Gamal El-Din, 2004), and manyother industrial chemicals (Legrini et al., 1993;Andreozzi et al., 1999; Alvares et al., 2001). In addition,

ozone treatment is often employed for pathogenicmicroorganism reduction (i.e., disinfection) (Loeb,2002; Paraskeva and Graham, 2002). Some of theadvantages of these processes include complete miner-alization of organic contaminants, production of lessharmful and more biodegradable by-products, and abil-ity to handle fluctuating flow rates and compositions(Zhou and Smith, 2001). The performance of AOP isalso affected by the presence of other water and waste-water constituents, such as natural organic matter,dissolved and suspended solids, and alkalinity, as wellas by water pH and temperature (Oppenlander, 2003).For example, suspended solids and color can hinderphotochemical reactions by light scattering and absorp-tion and may impair the performance of photochemicalAOPs, such as O3/UV, H2O2/UV, photo-Fenton, andTiO2/hn processes. Carbonate, bicarbonate, and chlor-ide ions, as well as some natural organic compounds areknown to act as radical scavengers. These compoundscompete with target pollutants for hydroxyl radicals;therefore their presence increases oxidant demands andlowers the treatment efficiency. In addition, the costs ofmaterials and equipment, as well as energy requirementsand efficiency must be taken into account when asses-sing the overall performance of AOPs (Legrini et al.,1993; Oppenlander, 2003).

DEGRADATION OF PHARMACEUTICALS INAQUEOUS SOLUTION BY OZONATION ANDADVANCED OXIDATION PROCESSES

Degradation of recalcitrant organic pollutants such aspharmaceuticals in water and wastewater can be achievedusing ozonation or AOPs. These treatment processes caneither eliminate such pollutants completely throughmineralization or convert them to the products that areless harmful to human health and the aquatic environ-ment. As the partial degradation of organic compoundsgenerally enhance their biodegradability (Alvares et al.,2001), the latter approach can be also used as a pretreat-ment of wastewaters containing pharmaceuticals prior tobiological treatment such as activated sludge. However,this approach may not be appropriate in the cases whereother organic matter is predominantly present, which isoften the case of municipal wastewater, because oxidantrequirement can be exceedingly high in order to achieveeffective degradation of trace target pollutants. Thus, it islikely more feasible to install ozonation or an AOP astertiary treatment after biological (secondary) treatment(see Figure 2), unless urinary sepacation is practiced assuggested by Larsen et al. (2004). In this case, the level oftreatment depends on the destination of treated effluent.Partial treatment may be sufficient if the effluent will bedischarged into surface water, although it is still neededto eliminate aquatic toxicity and improve biodegradabil-ity. However, if the effluent will supplement drinking

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water sources, which is increasingly common in semi-aridareas, complete removal or destruction of pharmaceuti-cals may be desired. As shown in Figure 2, ozonation andAOPs can be applied to water treatment as a part of pre-oxidation and/or disinfection steps. Wastewater frompharmaceutical production/formulation plants and hos-pital wastewater can be treated in a similar manner todomestic wastewater, with appropriate pretreatment suchas pH adjustment.

During the treatment by ozonation or advanced oxida-tion, organic pollutants such as pharmaceutical com-pounds undergo a series of oxidation and spontaneoustransformations. In other words, primary degradationproducts are often degradated further in prolonged treat-ment. In some cases, disappearance of parent pharmaceu-tical compounds does not indicate successful treatmentbecause the degraded products may be as biologicallyactive as the parent compounds. Likewise, some conven-tional water quality parameters can be used to evaluatethe effectiveness of the process, such as total organiccarbon (TOC), chemical oxygen demand (COD), dis-solved organic carbon (DOC), absorbable organic halo-gene (AOX), and aromaticity; however, they do notprovide direct information about the identity of degrada-tion products and the safety of treated water.Consequently, the analysis of pharmaceutical compoundsand their degradation products (by-products) using gaschromatography (GC) or liquid chromatography (LC)coupled with mass spectrometry (MS) has been increas-ingly common and becoming an asset in recent studies(Ternes, 2001; Debska et al., 2004).

In some cases, the rate constants for the reactionbetween a pharmaceutical compound and oxidant (i.e.,molecular ozone and hydroxyl radicals) has been deter-mined. These rate constants indicate the reactivity of the

compounds toward ozonation and AOPs and can be use-ful to model the treatment process (Beltran, 2003;Oppenlander, 2003). Typically, the reaction kinetics canbe described by a second order rate law:

r �Mð Þ ¼ d M½ �t=dt ¼ koxidant;M M½ � oxidant½ �

where r(�M) represents the rate of diminution of theconcentration of M (organic substrate=pharmaceuti-cals), and koxidant,M is the second order rate constant forthe reaction of oxidant (O3 or

�OH) and M. In addition tomolecular ozone and hydroxyl radical reactions, directphotolysis also contribute to the degradation of pharma-ceutical compounds during the AOPs in which UV-Visirradiation is involved. In such a case, the quantum yield(F), defined by the amount of reactant consumed peramount of photon absorbed, of the direct photolysismay be determined. These three types of reactions canoccur simultaneously and competitively, depending onthe AOP combination used, the presence of other waterand wastewater constituents, and pH.

A tabulated summary is presented in the Appendixwith major reaction conditions such as initial pH andtemperature. It should be noted that in most of the ozo-nation studies, only the amount of applied ozone hasbeen given, and there was not enough informationabout the actual amount of ozone utilized during thetreatment.

Antibiotics

Antibiotics are a group of pharmaceuticals used for thetreatment of both human and animals with bacterial andfungal infections. Many of the antibiotics are derivedfrom wholly or partially from certain microorganisms,but some are synthetic (e.g., sulfonamides). A wide

FIGURE 2. Possible points to apply ozonation and AOPs for the degradation of pharmaceuticals.

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range of antibiotics with diverse structures have beenused and subsequently found in the environment.Removal of some of these compounds has been studiedusing ozonation and AOPs (Table 1). In this review, theywill be divided into four sub-classes, including b-lactam,macrolide, quinolone, and tetracycline, as well as severalother types of antibiotics that do not belong to thesesubclasses. The chemical structure and molecular weightof these antibiotics are shown in Figures 3–7.

b-Lactam Antibiotics. b-Lactam antibiotics coveredin this review include amoxicillin, cefradine (cephradine),ceftriaxone, penicillins G and VK, and sultamicillin (seeFigure 3). These b-lactam antibiotics have been originallyisolated from molds, such as Penicillium chrysogenum andCephalosporium acremonium, and been modified to gaindifferent physico-chemical and pharmacological proper-ties (Merck & Co., 1999). These antibiotics generally havea property to inhibit bacterial cell wall synthesis. Amongthe b-lactam antibiotics here, cefradine and ceftriaxoneare the members of cephalosporins. It is known that mostof these penicillins and cephalosporins are excretedmainly in urine without major metabolic modifications(Merck & Co., 1999). Degradation of two intermediatesof cephalosporin antibiotics, including 5-methyl-1,3,4-thiadiazole-2-thiol (MMTD) and 5-methyl-1,3,4-thiadia-zole-2-methylthiol (MMTD-Me), have been also studiedand are included here.

Penicillins (including amoxicillin and sultamicillin).Ozonation and advanced oxidation treatment of variouspenicillin formulation effluents containing one of fourpenicillin-type b-lactam antibiotics, including penicillinsG and VK, amoxicillin, and sultamicillin has been exten-sively studied to improve the biodegradability of thesewastewaters. Akmehmet Balcioglu and Otker (2003)applied ozonation to synthetic penicillin VK formulationwastewater. About 70%, 40%, and 30% of initial COD(450 mg/L), TOC (162 mg/L), and aromaticity (0.456;absorbance at 254 nm) were removed by ozonation inone hour at an applied ozone dose of 2.96 g/L at pH 7or 11 at 20 �C. Biodegradability of the penicillin solution,measured by a ratio of 5-day biochemical oxygen demand(BOD5) and COD, was also improved considerably bythe ozone treatment from zero to 0.25. Removal of CODwas less at low pH, suggesting the importance of hydroxylradical reactions. Addition of 20 mM H2O2 (O3:H2O2

molar ratio=3.08:1) improved the COD removal to95% at pH 7 (Akmehmet Balcio�glu and Otker, 2003).

The feasibility of a number of AOPs was tested for thedegradation of penicillin formulation wash water (waste-water) containing amoxicillin trihydrate and clavulanicacid, a b-lactamase inhibitor (Arslan-Alaton andDogruel, 2004). The wastewater also contained otherinert substances such as additives, buffer, and flavors.The concentrations of these active and inert ingredientswere unknown ([TOC]0=920 mg/L, [COD]0=1,395 mg/L, alkalinity=85 mg CaCO3/L, pH 6.95). Among the

AOPs tested, ozonation at pH 11.5 and photo-Fenton/photo-Fenton-like process were more promising than theothers, namely H2O2/UV, Fenton, and Fenton-like pro-cesses (see Appendix). In addition to these AOPs, directphotolysis was also studied by Arslan-Alaton andDogruel (2004); however, it was virtually ineffective.About 50% of COD and 52% of TOC were removedfrom the wastewater by ozonation at an applied ozonedose of 2.76 g/L in 1 h at pH 11.5 (initial value). The pHvalue decreased presumably due to the formation oforganic acids during the ozonation, which inhibitedfurther COD removal. It was also noted that the CODremoval by ozonation was improved from 49% to 86%using phosphate buffer at pH 11.5. Photo-Fenton andphoto-Fenton-like processes [2 mM Fe2+ or Fe3+, 20mM H2O2, a 21-W low pressure Hg lamp, l=253.7 nm(1.73 · 10�4 Einstein�L�1�s�1)] are equally effective inCOD and TOC removals: 56 and 66% COD removaland 51 and 42% TOC removal, respectively. Completeconversion of 400 mg/L amoxicillin was also confirmedusing ozonation at pH 11.5 and photo-Fenton process(Arslan-Alaton and Dogruel, 2004).

Biodegradability of the penicillin formulation (amox-icillin + clavulanic acid) wastewater was improvedslightly from zero to 0.08, based on the BOD5/CODratio, by ozonation (Arslan-Alaton and Dogruel, 2004).Arslan-Alaton et al. (2004) investigated the biodegrad-ability improvement of a similar penicillin formulationwastewater containing amoxicillin and clavulanic acid asactive ingredients ([COD]0=830 mg/L) by ozonationand O3/H2O2 process. Under optimized conditions, thewastewater biodegradability was improved from 0 to 0.37and 0.45 by ozonation (2.5 g/L applied O3, pH 12) andO3/H2O2 process [2.5 g/L applied O3 (1.34 g/L consumed)and 2 mM H2O2, pH 10.5], respectively, in 1 h.Combination of chemical pre-oxidation and biologicaltreatment [0.23 mg COD/(mg mixed liquor suspendedsolids · day)] yielded an overall COD removal of 81%and 72% with optimized ozonation and O3/H2O2 process,respectively (Arslan-Alaton et al., 2004). A similar ozona-tion-biological treatment study has been reported on adifferent type of penicillin-type antibiotics, sultamicillin(Cokgor et al., 2004).

A series of studies has been reported concerning thetreatment of another type of penicillin formulation, peni-cillin G with procaine, a local anesthetic agent, by ozona-tion and photo-Fenton-like process. Degradation of thispenicillin formulation in synthetic wastewater was quan-tified by COD and TOC reductions in both cases. Arslan-Alaton and Caglayan (2005) demonstrated the enhance-ment of COD removal with increasing pH from 7 to 12and hence the important contribution of hydroxyl radicalreactions in the degradation of penicillin G-procaine for-mulation. Ozone absorption efficiency also increased atelevated pH. About 50% of initial COD (=600 mg/L)and TOC (=226 mg/L) and was removed by ozonation

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at pH 12 with an applied ozone dose of 1.8 g/L in 1 h.Overall second-order rate constants for the COD removalfrom the penicillin formulation by ozonation was deter-mined as 0.042 and 0.67 M�1�s�1 at pH 3 and 7, respec-tively (Arslan-Alaton and Caglayan, 2005).

Arslan-Alaton and Gurses (2004) investigated thephoto-Fenton-like process for the treatment of penicillinG-procaine formulation using a 125-W black light emittingUV-A in the range of 300 to 370 nm with an incident lightflux of 1.7 · 10�3 Einstein�min�1. Under optimum reactionconditions ([Fe3+]0=1.5 mM, [H2O2]0=25 mM, pH 3),56% and 42% of initial COD and TOC, respectively, wereremoved after 30 min of photo-Fenton-like treatment. Itwas also demonstrated that Fenton-like treatment withoutUV-A irradiation was slightly less effective in CODremoval from the penicillin formulation and thatneither UV-A alone, ferric ion with UV-A, nor H2O2/UV-A was effective (Arslan-Alaton and Gurses, 2004).Biodegradability of the penicillin G-procaine formulation

was improved from 0.25 to 0.45, while acute toxicity(assessed by Daphnia magna mortality) was reduced mark-edly after the 30 min of photo-Fenton-like treatment men-tioned previously (Arslan-Alaton and Gurses, 2004).

Andreozzi et al. (2005) investigated the kinetics andpathway of amoxicillin degradation by ozonation. Rateof reaction between amoxicillin and molecular ozone wasfound to be strongly pH dependent: from 4 · 103 M�1�s�1at pH 2.5 to 6 · 106 M�1�s�1 at pH 7. The second-orderrate constant for hydroxyl radical reaction with amoxi-cillin was also calculated as 3.93 · 109 M�1�s�1 at pH 5.5using H2O2/UV AOP to generate hydroxyl radicals(Andreozzi et al., 2005). Whereas hydroxylation of thephenol ring was confirmed, no evidence was found forS-oxidation of amoxicillin molecules by molecular ozone.No further degradation was noticed, probably due to theshort reaction time (up to 4 min) and the absence ofhydroxyl radicals in the system as 2-methyl-2-propanolwas used as a scavenger.

N

SHN

O

NH2

HO O

H

OHO

amoxicillin (365.41)

N

SHN

OO

H

OHO

penicillin G (334.39)

N

SHN

OO

O

H

OOpenicillin VK (388.47) K+

NH2

O

HN

NO

S

O

H

cefradine (349.40)

N

S

NO

O

HN

NO

S

O

S

NN

N

OH

OH2N

ceftriaxone (554.57)

NN

SHS

5-methyl-1,3,4-thiadiazole-2-thiol(MMTD; 132.20)

NN

SS

5-methyl-1,3,4-thiadiazole-2-methylthiol(MMTD-Me; 146.24)

SNH

OH2N

ON

OO

O

O N

O

SO

sultamicillin (594.65)

H

H

H

O

OH

OH

FIGURE 3. b-Lactam antibiotics and intermediates.

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Cefradine. Degradation of cefradine, a cephalos-porin-type b-lactam antibiotic has been studied using animmobilized titanium dioxide photo catalyst (Fan et al.,2002). Near-complete conversion of 0.2 mM (70 mg/L)cefradine was achieved by the TiO2/hn process using a 30-W UV lamp (l=253 nm) in 2 h (pH unknown) withcontinuous aeration. Decomposition of aromatic ring,amino and carboxyl groups was confirmed using UVand infrared spectroscopy. Addition of hydrogen perox-ide (1 to 5 g/L) in the system accelerated the cefradinedegradation, presumably due to enhanced production ofhydroxyl radicals through the photodecomposition ofhydrogen peroxide (Fan et al., 2002). Degradation by-product was not determined in this study.

Ceftriaxone. Similar to the case of penicillin VK, abuffered synthetic wastewater containing ceftriaxone wastreated by ozonation (Akmehmet Balcio�glu and Otker,2003). After 1 h of ozonation at an applied ozone dose of2.96 g/L, 74% of initial COD (=450 mg/L) and 50% of

initial TOC (=167 mg/L) was removed at pH 7. TheCOD reduction was slightly improved to 82% by elevat-ing pH from 7 to 11. Addition of hydrogen peroxide alsoimproved the COD removal. The absorbance at 254 nmin the synthetic wastewater was reduced by more than90%, indicating effective destruction of aromatic rings inthe ceftriaxone molecules. Biodegradability (i.e., the ratioof BOD5 and COD) of ceftriaxone solution was modestlyimproved from zero to 0.1 after the ozonation treatment(Akmehmet Balcio�glu and Otker, 2003). No degradationby-product was identified for the ceftriaxone ozonationor AOP.

MMTD and MMTD-Me. Photochemical degradationof two intermediates used for the synthesis of cefazolin,a cephalosporin-type b-lactam antibiotic, including 5-methyl-1,3,4-thiadiazole-2-thiol (MMTD) and 2-methyl-5-(methylthio)-1,3,4-thiadiazole (MMTD-Me),was reported (Bozzi et al., 2002; Lopez et al., 2002,2003). Lopez et al. (2002) demonstrated complete con-version of MMTD in less than 20 min by H2O2/UVtreatment ([H2O2]0=26 mg/L, a 17-W low-pressure Hglamp emitting at 185 and 254 nm with light intensity of2.8 · 10�6 Einstein�s�1, at 2�C), which was more effec-tive than UV irradiation alone. The quantum yield forMMTD photolysis was determined as 12mmol�Einstein�1. About 60% of initial TOC wasremoved during a prolonged H2O2/UV treatment of 1mg/L MMTD for 4 h, and about 90% of sulfur and14% of nitrogen was recovered as sulfate and nitrateions, respectively. Seven degradation intermediates and

O

O

O

OH

O

HO OH

O

HON

O

OOH

N

azithromycin (748.88)

O

O

O

OH

O

HO O

O

O

HON

O

OOH

clarithromycin (747.95)

O

O

O

OH

O

HO OH

O

O

HON

O

OOH

erythromycin (733.93)

O

O

O

OH

O

HO OH

N

O

HON

O

OOH

OOO

roxithromycin (837.05)

O

HOOH

OH

S

HO

NH

ON

lincomycin (406.54)

FIGURE 4. Macrolide antibiotics.

NNN

FO

OH

enrofloxacin (359.39)

N

O

ON

F

N

O

HO

ofloxacin (levofloxacin, 361.37)

O

FIGURE 5. Quinolone antibiotics.

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by-products were identified by LC-MS, and degradationpathway was proposed as shown in Figure 8 (Lopez et al.,2002). It is apparent that the 2-thiol group and the sulfuratom on the thiadiazole ring of MMTD are the primetargets of hydroxyl radical attack. The second-order rateconstant for MMTD degradation by H2O2/UV process

(hydroxyl radicals) was reported as 1.6 · 1010 M�1�s�1 at25�C (pH unknown) (Lopez et al., 2003).

Degradation of MMTD-Me, which is a 2-S-methyl deri-vative of MMTD, has been studied in a similar manner tothe latter compound (Bozzi et al., 2002; Lopez et al., 2003). Itwas noted thatMMTD-Mewasmore reactive toward directphotolysis presumably due to the higher extinction coeffi-cient (4970 M�1�cm�1) of MMTD-Me as compared withthat of MMTD (2100 M�1�cm�1), while it is less reactivetoward H2O2/UV treatment because of the electron with-drawal property of methyl group (Bozzi et al., 2002).Nevertheless, MMTD-Me could be completely convertedin 20 min by the equivalent H2O2/UV treatment mentionedabove at 25 �C (pH unknown). About 80% of initial TOCwas removed in 4 h, and a near-quantitative recovery ofsulfur as sulfate, as well as a 16% recovery of nitrogen asnitrate, was observed as well. Degradation intermedi-ates and by-products of the MMTD-Me photochemicaltreatment were also identified (Bozzi et al., 2002),which are shown in Figure 9. Unlike MMTD degrada-tion, demethylation and S-oxidation of 2-methylthiogroup likely precede the ring opening and mineraliza-tion of MMTD-Me, but no dimerizaiton (disulfide for-mation) may occur because of the presence of methylgroup. The quantum yield and second-order rate con-stant for MMTD-Me photolysis and H2O2/UV

N

N

NH2

H2NO

O

O

trimethoprim (290.32)

O

NHHO

HN

OH

O

O

O

H

OH

spectinomycin (332.35)

N

NO

O

NHN O

O

carbadox (262.22)

OH O OH O O

NH2

OHN

OH

HO

tetracycline (444.44)

FIGURE 7. Miscellaneous antibiotics.

H2N SO

O

HN

NNCl

sulfachlorpyridazine (284.72)

H2N SO

O

HN

N

N

sulfadiazine (250.27)

H2N SO

O

HN N

NO

O

sulfadimethoxine (310.33)

H2N SO

O

HN

N

N

sulfamerazine (264.30)

H2N SO

O

HN

N

N

sulfamethazine (277.34)

N O

HNS

O

OH2N

sulfamethoxazole (253.28)

H2N SO

O

HN

N

sulfapyridine (249.29)

N

SHNS

O

OH2N

sulfathiazole (255.31)

O NHNS

O

OH2N

sulfisoxazole (267.30)

N

OHNS

O

OH2N

sulfamoxole (267.30)

N N

SHNS

O

OH2N

sulfamethizole (270.32)

FIGURE 6. Sulfonamide antibiotics.

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treatment were reported as 12 mmol�Einstein�1 and 8.3· 108 M�1�s�1, respectively (Lopez et al., 2003).

Macrolide Antibiotics

Among the antibiotics shown in Figure 4, azithro-mycin, clarithromycin, erythromycin, and roxithromy-cin are macrolide antibiotics, and lincomycin is alincosamide antibiotic, which has an antibacterial spec-trum similar to macrolides. Although not shown inFigure 4, there is another lincosamide antibiotic calledclindamycin, which is a derivative of lincomycin andmore commonly prescribed nowadays because of itsimproved absorption property after oral administration(Merck & Co., 1999). These antibiotics are character-ized by a property to inhibit bacterial protein synthesis.These antibiotics are mostly excreted in the bile (Merck& Co., 1999).

Azithromycin. Huber et al. (2005) reported completedisappearance of azithromycin (initial concentration notspecified) in biologically treated municipal wastewater byozonation at an applied ozone dose of 3.5 mg/L at pH 7.They suggested that two tertiary-amine nitrogens in azithro-mycin molecule were the primary sites for molecular ozoneattack (see Figure 4 for the molecular structure of azithro-mycin). No degradation by-products were identified.

Clarithromycin and Erythromycin. Elimination(i.e., conversion) of trace clarithromycin and erythro-mycin in biologically treated municipal wastewaterhas been studied using ozonation recently. Ternes etal. (2003) reported an elimination of 0.21 mg/L clari-thromycin and 0.62 mg/L erythromycin to levelsbelow the limit of quantification with as low as 5mg/L of applied ozone at pH 7.2. A similar resultwas also reported by Huber et al. (2005). A biologi-cally treated municipal wastewater was spiked withclarithromycin and dehydroerythromycin (both at 2mg/L; see Figure 10 for the structure of dehydroery-thromycin, an environmental metabolite of erythro-mycin), along with other pharmaceuticals and muskfragrances, and was treated by ozonation at pH 7. Itwas found that an applied ozone dose of 3.5 mg/Lwas sufficient to eliminate these two macrolide anti-biotics. It was proposed a tertiary-amine nitrogen inboth macrolides and a carbon-carbon double bond indehydroerythromycin molecule to be ozone attacksites (Huber et al., 2005).

Roxithromycin. It has been reported that the reactiv-ity of roxithromycin toward ozonation is similar to theother macrolide antibiotics discussed above (Ternes et al.,2003; Huber et al., 2005). In addition, one kinetic study ofroxithromycin ozonation has been reported (Huber et al.,2003). It was shown that this macrolide antibiotic reactedwith ozone rapidly (kO3 > 105 at pH 7, 20 �C) and that therate constant for the direct ozone reaction was pH-depen-dent. The rate constant reaches to its maximum value of4.5 · 106 M�1�s�1 at pH 8.8, which corresponds to the pKa

of roxithromycin, indicating the tertiary-amine nitrogen isthe primary site of ozone attack (Huber et al., 2003).

N N

S SH

hν

N N

SN N

S SO3H

•OH

N N

S S S NNS

N N

S S NN

S

- S

N N

S O

or

SO3H

N N

S SO3H

O

N N

S S NN

S

hν

•OH

•OH

- S

- SO4-

N N

S

O

•OH

ring opening and

mineralization

•OH•OH

•OH

MMTD

OOOO

FIGURE 8. Proposed pathway for MMTD degradation by H2O2/

UV process (Lopez et al., 2002).

N N

S S hν

N N

S SH

MMTD-Me

N N

S S

O•OH

N N

Shν

•OH

•OH

•OH

mineralization

FIGURE 9. Degradation of MMTD-Me by H2O2/UV process

(Bozzi et al., 2002).

O

O

OOH

O

HO OH

O

HON

O

OOH

O

FIGURE 10. Chemical structure of dehydroerythromycin, an

environmental metabolite of erythromycin (Huber et al., 2005).

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Lincomycin. Qiang et al. (2004) investigated a rapidreaction of lincomycin toward ozonation (kO3 > 105

M�1�s�1 at pH >7) using a stopped-flow technique.Similar to the case of roxithromycin, lincomycin degrada-tion by ozonation is pH dependent (pKa of pyrrolidinenitrogen=7.8), and the contribution of hydroxyl radicalreactions is considered to be negligible. Besides the pyr-rolidine nitrogen, a sulfur atom of methylthio group issuggested as a prime target of ozone attack (see Figure 11).Based on their kinetic model, the absolute second-orderrate constants for lincomycin ozonation were calculatedas 3.26 · 105 and 2.43 · 106 M�1�s�1 for protonated andneutral forms, respectively (Qiang et al., 2004).

In addition to ozonation, Addamo et al. (2005)reported photo-catalytic degradation of lincomycinusing two types of TiO2 catalysts. Addition of TiO2 pow-der (0.4 g/L Degussa P25 or 1 g/L Merck TiO2) stronglyenhanced the degradation of lincomycin from UV irradia-tion alone with a 125 W medium pressure Hg lamp with aphoto flux of 8.5 mW�m�2. Complete conversion of 50mg/L lincomycin was achieved in 2 h by the TiO2/hntreatment at pH 6 and 20 �C, and about 60% of initialTOC was removed from the solution. Although S-oxida-tion of methylthio group and oxidation of pyrrolidinering followed by a cleavage of amide bond were suggestedas possible routes of lincomycin degradation by TiO2

process (Addamo et al., 2005), the actual degradationintermediates and by-products were not determined,with the exception of sulfate ion.

Quinolone Antibiotics

Two types of quinolone antibiotics, including enro-floxacin and ofloxacin, have been studied for theirdegradation by ozonation and AOPs. More specifi-cally, both are fluoroquinolones. While enrofloxacin isused as a veterinary antibiotic, ofloxacin is intendedfor human consumption. These compounds have aproperty to inhibit the activity of bacterial DNA gyr-ase (Merck & Co., 1999). The chemical structure andmolecular weight of these quinolone antibiotics areshown in Figure 5. Please note that ofloxacin is aracemic mixture of chiral molecules, and the biologi-cally active S-enantiomer, which is also called levoflox-acin, is shown in the figure. It is known that the

quinolones are variably metabolized in the liver andexcreted in the urine (Merck & Co., 1999).

Enrofloxacin. Akmehmet Balcioglu and Otker (2003)demonstrated that ozone treatment was effective indegrading enrofloxacin in synthetic wastewater. About90% of initial COD (=450 mg/L) and 50% of initialTOC (165 mg/L) was removed by ozonation in onehour at an applied ozone dose of 2.96 g/L and pH 7.The COD removal was less at pH 3 and pH 11, 65% and79%, respectively. Addition of hydrogen peroxide (10 to100 mM) did not show any impact on the COD removal.Biodegradability, as expressed by the ratio of BOD5 andCOD, of the synthetic wastewater was improved consid-erably from 0.07 to 0.38 after the ozone treatment at pH 7(Akmehmet Balcio�glu and Otker, 2003). The same groupof authors also investigated the use of negatively chargedzeolite as natural adsorbent for the treatment of enroflox-acin, which has three tertiary-amine nitrogens in its mole-cular structure, and subsequent decontamination andregeneration of zeolite by ozonation (Otker andAkmehmet-Balcio�glu, 2005). They demonstrated the suc-cessful decontamination of enrofloxacin-loaded zeolite(equilibrated with 200 mg/L enrofloxacin for 24 h) by30 min of ozonation (consumed ozone=100 mg), sug-gesting that the zeolite adsorption-ozonation systemmight be a good method to control antibiotic pollutiondue to animal manure land application.

Ofloxacin. Andreozzi et al. (2004) investigated theozonation, H2O2/UV, and TiO2 photo-catalysis for thedecontamination of an aqueous pharmaceutical mixturecontaining ofloxacin, as well as five other drugs, namelycarbamazepine, propranolol, clofibric acid, diclofenac,and sulfamethoxazole. It was shown that 2 min of ozona-tion (13.9 mg/L absorbed O3) as well as H2O2/UV treat-ment ([H2O2]0=5–10 mM, a low-pressure Hg lamp, at254 nm, 2.51 · 10�6 Einstein�s�1) was sufficient for thecomplete conversion of 560 mg/L ofloxacin at pH 7.4 and25 �C, whereas TiO2/hn process was less effective. Algal(Synechococcus leopoliensis) and protozoan (Brachionuscalyciflorus) toxicities of the drug mixture were eliminatedcompletely by the ozone treatment in 2 min. The H2O2/UV treatment was also effective for algal toxicity reduc-tion; however, it was less effective for protozoan toxicityreduction. Finally, the TiO2/hn treatment was ineffectivein either case (Andreozzi et al., 2004), probably becauseof insufficient drug conversion. Degradation by-productsof ofloxacin were not determined.

Sulfonamide Antibiotics

Sulfonamides are synthetic antibiotics also called sulfadrugs that inhibit multiplication of bacteria by acting ascompetitive inhibitors of p-aminobenzoic acid in the folicacid metabolism cycle (Merck & Co., 1999). A variety ofsulfonamides have been produced, consumed, and subse-quently detected in the environment, and some of them

O

HOOH

OH

S

HO

NH

ON

O3 or •OHO3 or •OH

•OH

FIGURE 11. Possible sites of molecular ozone and hydroxyl

radical attacks in a lincomycin molecule (Qiang et al., 2004;

Addamo et al., 2005).

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have been studied for their degradation by ozonation andAOPs. These compounds include sulfachlopyridazine, sul-fadiazine, sulfadimethoxine, sulfamerazine, sulfametha-zine, sulfamethizole, sulfamethoxazole, sulfamoxole,sulfapyridine, sulfathiazole, and sulfisoxazole. They havea common core chemical structure (p-aminobenzene sulfo-namide), as can be seen in Figure 6. These sulfonamideantibiotics are metabolized mainly by the liver to acetylatedforms (at p-amino group) and glucuronides, and subse-quently excreted in the urine (Merck & Co., 1999).

Sulfachlorpyridazine. Adams et al. (2002) reportedthe rapid conversion of sulfachlorpyridazine by ozona-tion in a pre-filtered river water sample spiked with thisand three other antibiotics at an initial concentration of50 mg/L each. More than 95% of initial sulfachlorpyrida-zine was converted by ozonation in 1.3 min with a utilizedozone dose of 0.3 mg/L at pH 7.5.

The second order rate constant for the reaction betweenhydroxyl radicals and sulfachlorpyridazine, as well as thosefor nine other sulfonamide antibiotics, has been reported(Boreen et al., 2004; 2005) as shown in Table 2. Fentonprocess was employed to generate hydroxyl radicals, and acompetition kinetics method was used to estimate the rateconstants using acetophenone as a reference compound.Unlike direct photolysis, which is strongly affected by thetypes of heteroaromatic ring attached to sulfonamide nitro-gen, the rate constants for hydroxyl radical reactions arenearly equal and close to diffusion-controlled limit of 1010

M�1�s�1. It has been also shown that the reactivity of thesesulfonamides toward direct photolysis is dependent on theprotonation state of the concerned compound (Figure 12,Table 2; Boreen et al., 2004; 2005), although the effect ofpH on hydroxyl radical reactions has not been studied. Nodegradation by-product/intermediate was identified for sul-fachlorpyridazine ozonation or AOP.

Sulfadiazine. Degradation of sulfadiazine has beenstudied by ozonation (Huber et al., 2005) and TiO2 photo-catalysis (Calza et al., 2004a). In addition, as shown inTable 2, the second-order rate constant for the reactionbetween hydroxyl radical and this sulfonamide antibiotichas been reported (Boreen et al., 2005). Sulfadiazine wascompletely converted by ozone (2 mg/L applied ozonedose, pH 7) in a biologically treated wastewater samplespike with this antibiotic at a concentration of 2 mg/L,along with several other sulfonamide antibiotics, macrolideantibiotics, estrogens, and iodinated X-ray contrast media(Huber et al., 2005). Calza et al. (2004a) demonstratedpartial mineralization of 15 mg/L sulfadiazine by TiO2

photocatalysis. About 80% of sulfur atom was recoveredas sulfate after 30 min of treatment with 200 mg/L TiO2

and UV irradiation using a 1500-W xenon lamp (1.35 ·10�5 Einstein�min�1) and a 340 nm cut off filter at 50 �C(pH unknown). On the other hand, the yield of inorganicnitrogen (i.e., ammonium ion) was about 15%, indicatingthe formation of stable organic by-products bearing nitro-gen atom(s), although these by-products were not deter-mined. During the TiO2/hn treatment, formation ofhydroxylated intermediate of sulfadiazine was suggested(Figure 13), which quickly disappeared as the degradationproceeded (Calza et al., 2004a). As aniline is not particu-larly persistent toward advanced oxidation (Sauleda andBrillas, 2001; Li et al., 2003), the N-bearing intermediatessuggested by Calza et al. are likely derived from 2-amino-pyrimidine (R-NH2 in Figure 13) released by the cleavageof S-N bond of sulfadiazine.

Sulfadimethoxine, Sulfamerazine and Sulfathiazole.Similar to the cases of other sulfonamide antibiotics,rapid conversion of sulfadimethoxine, sulfamerazine,and sulfathiazole by ozonation in river water wasreported (Adams et al., 2002; Huber et al., 2005), and

TABLE 2. pKa, Second-Order Rate Constant for Hydroxyl Radical Reaction (k�OH), and Quantum Yield (F) of Sulfonamide Antibiotics (Boreen

et al., 2004; Boreen et al., 2005)

FP (mM�Eins�1)

pKa,1 pKa,2 k�OH (109 M�1�s�1) SHþ2 SH S�

Sulfachlorpyridazine 2 5.9 4.4 nd 3.0 · 10�4 2.3 · 10�3

Sulfadiazine 2 6.4 3.7 nd 4.0 · 10�4 1.2 · 10�3

Sulfadimethoxine 2.9 6.1 6.1 nd 1.0 · 10�5 4.0 · 10�4

Sulfamerazine 2.5 7 3.8 nd 2.3 · 10�4 3.0 · 10�3

Sulfamethazine 2.6 8 5.0 nd 3.0 · 10�4 5.0 · 10�3

Sulfamethizole 2.1 5.3 4.9 � 0.01 � 5.0 · 10�3 0.05Sulfamethoxazole 1.6 5.7 5.8 0 0.5 0.09Sulfamoxole nd 7.4 nd nd nd ndSulfathiazole 2.2 7.2 7.1 0.02 0.07 0.40Sulfisoxazole 1.5 5.0 6.6 0.7 0.17 0.07

Note: k�OH was determined for neutral sulfonamide. SHþ2 ; SH, and S� represent cationic, neutral, and anionic forms ofsulfonamides (see Figure 12). nd=not determined.

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comparable second-order rate constants for the hydroxylradical reactions of these sulfonamide antibiotics weredetermined (see Table 2; Boreen et al., 2005). Calzaet al. (2004a) demonstrated near-complete conversion ofthese three sulfonamides by TiO2/hn treatment (seeSulfadiazine for the reaction conditions). Like the caseof sulfadiazine, high levels (70% to 100%) of inorganicsulfur were recovered as sulfate after the TiO2/hn treat-ment of sulfadimethoxine, sulfamerazine, and sulfathia-zole, while the yields of inorganic nitrogen (as ammoniumand nitrate) were substantially different: 70%, 18%, and90%, respectively (Calza et al., 2004a). The behavior ofsulfamerazine resembles that of sulfadiazine, indicatingpossible formation of persistent nitrogenous by-products.On the other hand, the amount of ammonium ionreleased by sulfadimethoxine and sulfathiazole degrada-tion implies effective destruction of the heterocyclic ring,i.e., 2,6-dimethoxypyrimidine and thiazole, respectively.Several degradation intermediates were detected duringthe TiO2/hn treatment of these sulfonamide antibiotics(Calza et al., 2004a), which are shown in Figure 14. Itwas noted that 2-aminothiazole, a possible degradationintermediate of sulfathiazole, was not detected probablydue to its high reactivity toward hydroxyl radicals (Calzaet al., 2004a).

Sulfamethazine. Rapid conversion of sulfametha-zine by ozonation was reported (Adams et al., 2002).Ozonation at an absorbed ozone dose of 0.3 mg/L wassufficient to transform 50 mg/L sulfamethazine in a pre-filtered river water sample at pH 7.5 in 1.5 min. The

second-order rate constant for the reactions between sul-famethazine and hydroxyl radicals was also reported(Boreen et al., 2005) as shown in Table 2. It has beennoted a slower reaction of N4-acetylated metabolite ofthis sulfonamide antibiotic with ozone in a biologicallytreated wastewater sample (Huber et al., 2005). No degra-dation by-product/intermediate was identified.

Sulfamethoxazole. Like many other sulfonamideantibiotics, sulfamethoxazole is readily degradable byozonation. Ternes et al. (2003) demonstrated that as lowas 5 mg/L of applied ozone could completely eliminate(i.e., convert) 0.62 mg/L sulfamethoxazole present in abiologically treated municipal wastewater below its detec-tion limit. Similar results were also reported elsewhere(Huber et al., 2003, 2005). A negative effect of dissolvedorganic matter on the degradation of sulfamethoxazolewas noted during the ozonation of this sulfonamide

H3N SO

O

HN R

Ka,1H2N S

O

O

HN R H2N S

O

ON R

Ka,2

SH2+ SH S-

FIGURE 12. Protonation states of the sulfonamide antibiotics (Boreen et al., 2004). SHþ2 ; SH, and S� represent cationic, neutral, and anionic

forms, respectively.

H2N SO

O

HN R H2N S

O

O

HN R

•OH

•OH

OHparent sulfonamide

H2N SO

OOH + R-NH2

•OH

•OH

H2N SO

OOH

OH

R-NH2 +

SO42-

•OH

SO42-

Further degradation, release of NH 4+

hydroxylated sulfonamide

•OH •OH

FIGURE 13. Degradation of sulfonamides by advanced oxidation

process (Calza et al., 2004a).

H2N SO

O

HN N

NO

O

H2N SO

O

HN

N

N

hydroxylated sulfamerazine

OH

hydroxylated sulfadimethoxine(aniline side)

H2N SO

O

HN N

NO

OHO

hydroxylated sulfadimethoxine(pyrimidine side)

H2N NN

O

O

2,6-dimethoxy-4-aminopyrimidine

H2NN

N

OH

4-methyl-2-aminopyrimidine

N

SHNS

O

OH2N

hydroxylated sulfathiazole

OH

FIGURE 14. Possible degradation intermediates of sulfadi-

methoxine (two hydroxylated compounds, 2,6-dimethoxy-4-amino-pyrimidine) and sulfamerazine (one hydroxylated compound, 4-

methyl-2-aminopyrimidine) and sulfathiazole (one hydroxylated

compound) detected during TiO2/hn treatment (Calza et al.,

2004a). See Figure 13 for the proposed general degradationpathway.

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antibiotic in natural river water samples (Huber et al.,2003). Huber et al. (2003) also determined the second-order rate constants for both ozonation (2.5 · 106

M�1�s�1) and advanced oxidation (H2O2/UV; 5.5 ·109 M�1�s�1) of sulfamethoxazole at pH 7 and 25 �Cby a competition kinetics method with phenol and p-chlorobenzoic acid as reference compounds, respec-tively. The rate constant for the hydroxyl radical reac-tions is in close agreement with the one reported forneutral sulfamethoxazole by Boreen et al. (2004) usingFenton process to generate hydroxyl radicals (Table 2).Since majority of this sulfonamide antibiotic is presentin the anionic form at pH 7 (pKa2=5.7), it is apparentthat the dissociation of sulfonamide-N proton (seeFigure 12) have no effect on the advanced oxidationof this antibiotic. It was suggested that aniline nitrogenwas the primary target of molecular ozone attack inthe sulfamethoxazole molecule (see Figure 6) and thatprotonation of this nitrogen would diminish its reactiv-ity toward ozonation (Huber et al., 2003). The firstpoint is evident from the fact that N4-acetylated sulfa-methoxazole (Figure 15), which is a metabolite of thissulfonamide antibiotic (Merck & Co., 1999), is shownto be relatively persistent during ozonation treatmentas compared with the non-acetylated parent compound(Huber et al., 2005).

Andreozzi et al. (2004) reported the effective detoxifi-cation of a mixture of pharmaceuticals containing 2.24mg/L sulfamethoxazole and five other drugs by ozona-tion and H2O2/UV treatment (see Ofloxacin, quinoloneantibiotic for details). No degradation intermediate/by-product was determined for sulfamethoxazole ozonationand advanced oxidation.

Sulfamethizole, Sulfamoxole, and Sulfisoxazole.Boreen et al. (2004) determined the second-order rateconstants (k�OH) for sulfamethizole and sulfisoxazole(see Table 2 and Sulfachlorpyridazine) using the Fentonprocess to generate hydroxyl radicals. They also notedthat another sulfonamide antibiotic, sulfamoxole, was notstable in weakly acidic (pH 4) solution. As a result, nokinetic study was conducted on this compound (Boreenet al., 2004). No further study was found on the degrada-tion of these sulfonamide antibiotics by ozonation orAOP.

Sulfapyridine. Huber et al. (2005) reported near-complete conversion of 2 mg/L sulfapyridine by ozonation

(as low as 2 mg/L applied O3 at pH 7) in a biologicallytreated wastewater sample. No further study has beenreported on the ozonation or advanced oxidation of thissulfonamide antibiotic.

Other Antibiotics

In addition to the groups of antibiotics discussed here,several other types of antibiotics have been studied fortheir degradation by ozonation and/or AOPs (Figure 7).Carbadox is a veterinary antibiotic, which can be addedto swine as a growth promoter although its use has beenbanned in Europe because of its suspected carcinogenicityand mutagenicity (Hutchinson et al., 2005).Spectinomycin and tetracycline are bacteriostatic antibio-tics that inhibit bacterial protein synthesis by binding the30S subunit of the ribosome (Merck & Co., 1999).Trimethoprim is a chemotherapeutic antibiotic, which iscommonly used in combination with sulfamethoxazole, asulfonamide antibiotic to maximize its effect on inhibitingfolate synthesis of bacteria (Merck & Co., 1999).

Carbadox and Trimethoprim. Adams et al. (2002)demonstrated rapid conversion of carbadox and tri-methoprim by ozonation in a pre-filtered river watersample spiked with these antibiotics at an initial con-centration of 50 mg/L each. More than 95% of theseantibiotics were converted by ozonation for 1.5 minwith a utilized ozone dose of 0.3 mg/L at pH 7.5.Ternes et al. (2003) also reported a similar reactivityof trimethoprim originally present in biologically trea-ted wastewater toward ozonation. No further study hasbeen reported on ozonation or advanced oxidation ofthese antibiotics.

Spectinomycin. Quang et al. (2004) investigated thekinetics of spectinomycin ozonation using a stopped-flowtechnique. As can be seen in Figure 7, this antibiotic hastwo secondary amine groups (pKa=7.10 and 8.90) thatcan be primary targets of molecular ozone attacks. It wasshown that the protonation of these two amine groups inacidic solution (pH < 5) diminished the reactivity ofspectinomycin toward ozonation. Absolute rate constantwas determined as 1.27 · 106 M�1�s�1 and 3.30 · 105

M�1�s�1 for neutral and mono-protonated forms of spec-tinomycin, respectively (Qiang et al., 2004). No degrada-tion by-product/intermediate was determined in thisstudy.

Tetracycline. Effective degradation and partialmineralization of tetracycline by TiO2 photocatalysiswas reported (Di Paola et al., 2004; Addamo et al.,2005). Almost complete conversion of 50 mg/L tetracy-cline was achieved in two hour by the TiO2/hn treatmentwith 0.4 g/L of catalyst and a 125 W medium pressureHg lamp with a photon flux of 8.5 mW�cm�2 at 40 �C,and about 90% of TOC was removed in 6 h (Di Paolaet al., 2004). On the other hand, direct photolysis wasmuch less effective in tetracycline conversion and miner-alization (Addamo et al., 2005). While the formation

N OHNS

O

OHN

O

FIGURE 15. N4-acetylsulfamethoxazole, an environmentally rele-

vant metabolite of sulfamethoxazole (Huber et al., 2005).

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of 4a,12a-anhydro-4-oxo-4-dedimethylaminotetracyclinewas suggested as a photolytic by-product (Addamoet al., 2005), no degradation by-product/intermediatewas determined for the tetracycline decomposition cata-lyzed by TiO2.

Summary of Antibiotics

All antibiotics discussed here are fairly reactive towardozonation and advanced oxidation in aqueous media.Their high reactivity is accounted for the presence ofone or more reactive functional groups/moieties in theantibiotic molecules, such as amine nitrogen, sulfur, acti-vated aromatic ring, and carbon-carbon double bond. Itshould be noted that acetylated sulfonamide metabolites,such as N4-acetylsulfamethoxazole (Figure 15), are moreresistant to ozonation than the parent compounds (Huberet al., 2005) because the acetylation blocks the anilineamino group that is one of the targets of molecularozone reactions. It has been demonstrated in the studiesreviewed here that the complete conversion of antibioticsis readily achievable by either ozonation or AOP; how-ever, their mineralization is not always confirmed.Similarly, the identity and quantity of degradation by-products arising from ozonation and advanced oxidationtreatment are largely unknown for most antibiotics.Sulfonamides are the only class of antibiotics whosedegradation pathway has been proposed (Figure 13),although it applies only to AOPs and their degradationmechanisms by molecular ozone alone is still elusive. Onthe other hand, kinetic information for ozonation andadvanced oxidation is available for many antibiotics.The kinetic data consistently indicate the high reactivityof antibiotics toward molecular ozone (kO3 >105

M�1�s�1) and hydroxyl radical (k�OH >109 M�1�s�1).The effect of ozonation and AOPs on the toxicity andbiodegradability has been studied only for some b-lactamantibiotics, two quinolones, and one sulfonamide. Morestudies will be needed to ensure not only the completetransformation of antibiotics but also the diminution oftoxicity toward environmental microbes and other aqua-tic organisms, because once released these compoundscan directly interact with natural bacterial population.Although the concentrations of antibiotics in the environ-ment are generally low (less than one microgram perliter), the bacterial population and aquatic life may bedisturbed and resistance against particular types of anti-biotics may be developed by pathogenic bacteria in theenvironment (Kummerer, 2004). For the same reason,major degradation intermediates/byproducts of antibio-tics should be identified and characterized. Finally, thereare several other antibiotics that have been detected inwastewater effluent and the environment but not studiedfor their degradation by ozonation or AOP. These anti-biotics include ciprofloxacin (Miao et al., 2004), clinda-mycin (Christian et al., 2003; Batt and Aga, 2005),norfloxacin (Kolpin et al., 2002; Miao et al., 2004),

doxycycline (Miao et al., 2004), and oxytetracycline(Kolpin et al., 2002).

Anticonvulsants and Anti-anxiety Agents

Anticonvulsants, sometimes called antiepileptics, are thegroup of pharmaceuticals used in drug therapy to controlseizures. The anticonvulsants covered in this review includecarbamazepine, diazepam, and primidone. Some anticon-vulsants are also useful to treat some types of psychiatricdisorders such as anxiety, depression, and mood disorder.Therefore, an anti-anxiety agent, buspirone, is also includedin this section. Chemical structure and molecular weight ofthese pharmaceuticals are shown in Figure 16.

Buspirone. Calza et al. (2004b) investigated photo-catalytic transformation of buspirone, an anti-anxietyagent, using TiO2 as a catalyst to simulate the metabolicsystem of living organisms. Complete transformation of15 mg/L buspirone was achieved by TiO2/hn treatmentwith 200 mg/L TiO2 and a 1500-W xenon lamp with a 340nm cut-off filter (1.35 · 10�5 Einstein�min�1) at 50 �C,and a number of degradation intermediates were identi-fied by LC-MS-MS as shown in Figure 17. It was sug-gested that hydroxylation of pyrimidine ring andazaspirone decane dione substructure occurred simulta-neously followed by cleavage and eventual loss of pyri-midine ring. The formation of another intermediate, 1-pyrimidinyl-piperazine, was also confirmed (Figure 17).These intermediates subsequently disappeared within 30min of the TiO2/hn treatment (Calza et al., 2004b).

Carbamazepine. Carbamazepine is a carboxamide-type anticonvulsant widely used to control generalizedtonic-chronic seizures (Merck & Co., 1999). This pharma-ceutical has been found ubiquitously in the aquatic envir-onment often at 1 to 2 mg/L levels (Ternes, 1998; Sacheret al., 2001; Stackelberg et al., 2004) and is probably one ofthe most persistent pharmaceuticals in the environment.Carbamazepine is highly resistant to biodegradation(Clara et al., 2004). It has been also suggested that

N

O NH2

carbamazepine (236.27)

N

O

O

NNN

N

buspirone (385.51)

N N

Cl

O

diazepam (284.74)

NH

HN

O

O

primidone (218.25)

FIGURE 16. Anticonvulsants and anti-anxiety agents.

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carbamazepine-glucuronide conjugate can be cleaved insewage treatment plant, which may contribute to theincrease of environmental concentration of this drug(Ternes, 1998).

High reactivity of carbamazepine toward ozonationwas reported by several groups of researchers(Andreozzi et al., 2002; Ternes et al., 2002; Huber et al.,2003). For example, an applied ozone dose of 0.5 mg/Lwas sufficient to convert 1 mg/L carbamazepine spiked insurface water after flocculation ([DOC]=1.3 mg/L) atpH 7.8 (Ternes et al., 2002). Andreozzi et al. (2002) alsoreported a complete conversion of 118 mg/L carbamaze-pine and 30% mineralization monitored as CO2 evolutionin one hour of ozonation treatment at a utilized ozonedose of 1.0 mg/L at pH 5.5. Algal toxicity of carbamaze-pine was diminished after the ozone treatment (Andreozziet al., 2002).

Two considerably large second-order rate constantsfor the molecular ozone reaction of carbamazepine at 25�C were reported: 7.81 · 104 M�1�s�1 (Andreozzi et al.,2002) and 3 · 105 M�1�s�1 (Huber et al., 2003). The rateconstant was pH-independent (Huber et al., 2003), which

is apparent from the molecular structure of carbamaze-pine. A number of degradation intermediates were iden-tified in two separate studies (Andreozzi et al., 2002;McDowell et al., 2005) as shown in Figure 18. It wassuggested that an ozone molecule first attacked thenon-aromatic carbon-carbon double bond of carbamaze-pine, leading to ring opening through the Criegeemechanism, followed by ring closure to form the quinazo-line moiety of 1-(2-benzaldehyde)-4-hydro-(1H,3H)-quina-zoline-2-one (BQM) shown in the figure [see McDowell etal. (2005) for the detail of proposed reaction mechanism].Quinazoline derivatives were also detected in the surfacewater treated by ozone in a full scale water treatmentplant (McDowell et al., 2005). McDowell et al. (2005)also presented a kinetic model that accounted for theformation and subsequent degradation of theseintermediates.

Vogna et al. (2004b) reported a different pathway ofcarbamazepine degradation by H2O2/UVAOP (Figure 19),which involves epoxidation of the carbon-carbon doublebond followed by the formation of a heteroaromatic inter-mediate, acridine. Carbamazepine could be converted

N

O

O

NNN

N

buspirone

N

O

O

NNN

NN

O

O

NNN

N

NHNN

N

OH

HO

N

O

O

NNHN

H2N

N

O

O

NHN

N

O

O

NNN

N

OH

HO

N

O

O

NNHN

H2N

OH

N

O

O

NNN

N

O

•OH •OH

•OH

•OH

•OH

•OH•OH

•OH

•OH

oxo-buspirone

hydroxy-bispirone hydroxy-bispirone

1-pyrimidinyl piperazine

dihydroxy-buspirone

despyrimidinyl-hydroxy-buspirone-amidine

despyrimidinyl buspirone

despyrimidinyl buspirone-amidine

FIGURE 17. Oxidative transformation of buspirone by TiO2/hn treatment (Calza et al., 2004b).

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rapidly by H2O2/UV treatment; about 4.7 mg/L carbama-zepine was completely converted within 4 min of treatmentwith 170 mg/L H2O2 and UV irradiation at 254 nm using a17 W low-pressure Hg lamp (2.7 · 106 Einstein�s�1) at pH5.0 (Vogna et al., 2004b). About 35% of initial TOC wasalso removed during the 4-min H2O2/UV treatment, whichis apparently more efficient than ozonation (Andreozziet al., 2002). A negative effect of humic substances on theaqueous carbamazepine degradation was observed (Vognaet al., 2004b). Two different second-order rate constantsfor the hydroxyl radical reaction of carbamazepine werereported: 2.05 · 109M�1�s�1 (Vogna et al., 2004b) and 8.8 ·109 M�1�s�1 (Huber et al., 2003). The difference might beoriginated from the difference in the methods of rate con-stant determination: the former group used a direct modelfitting with the experimental data, whereas the latter groupemployed a competition kinetics method using p-chloro-benzoic acid as a reference compound.

Recently, the carbamazepine degradation by TiO2

photocatalysis was investigated in a series of study (Dolland Frimmel, 2004, 2005a, 2005b, 2005c). Doll andFrimmel (2004) showed that as compared with otherpersistent pharmaceuticals including clofibric acid, iome-prol, iopromide, carbamazepine could be rapidlydegraded by TiO2/hn process. More than 90% of initialcarbamazepine (4.2 mg/L) was converted in nine minutesby treatment using 100 mg/L TiO2 (P25) and a 1000-Wxenon short-arc lamp as a source of simulated solar UVrays (1.35 · 104 Einstein�m�2�s�1, l <400 nm) at pH 6.5(Doll and Frimmel, 2005a). A successful pilot-scale inves-tigation of the continuous treatment of carbamazepine bythe TiO2/hn process was reported using a cross-flowmicrofiltration to retain catalyst (Doll and Frimmel,2005b). The degradation of carbamazepine was stronglyinhibited in the presence of NOM in a lake water sampleby competing for hydroxyl radicals, as well as by deacti-vating TiO2 catalyst surface by adsorption (Doll and

Frimmel, 2005a). Doll and Frimmel (2005c) identified anumber of carbamazepine degradation products duringthe TiO2/hn treatment, and a degradation pathway simi-lar to the one proposed for H2O2/UV AOP (see Figure 19)was indicated. They also suggested the possible environ-mental impact of acridine and its derivatives generatedduring the carbamazepine degradation as these com-pounds are polycyclic heteroaromatics known to betoxic to aquatic organisms, especially via photosensitiza-tion (Wiegman et al., 2002).

Diazepam. Diazepam is a benzodiazepine-type anti-anxiety agent that is also used to treat many otherneurological and psychiatric disorders such as sleep dis-order, movement disorders, and motion sickness (Merck& Co., 1999). After administration, diazepam is meta-bolized and excreted mainly in the urine. Huber et al.(2003) reported very slow degradation of diazepam bymolecular ozone with a second-order rate constant of0.75 M�1�s�1 at 20�C. Ozonation of 142 mg/L diazepamin natural water samples resulted in 24% to 65% con-version of this pharmaceutical (applied ozone dose=2mg/L, 10 min, pH 8, 10�C). Their kinetic analysisrevealed that the diazepam was mostly degradedthrough hydroxyl radical reactions (k�OH=7.2 · 109

M�1�s�1) (Huber et al., 2003). Deactivation of aromaticrings by the imine group (�C=N-) and chlorine atom(Beltran, 2003) may be the reason for the low reactivityof diazepam toward ozonation. No degradation inter-mediate/by-product was determined.

Primidone. Primidone is an anticonvulsant of the pyr-imidinedione class, which used to treat disorders of move-ment such as tremor (Merck & Co., 1999). This drug ismetabolized in the liver to phenobarbital, which is also ananticonvulsant (Bergey, 2004), and excreted in the urine.Ternes et al. (2002) reported moderate reactivity of primi-done toward ozonation. About 87% of 1 mg/L primidonespiked in a flocculated surface water sample was converted

N

O

O

N

N

O NH2

O3 O3N

O

O

HN

O

O3NO

HN

O

OH

O

carbamazepine BQM BQD BaQD

(further degradation)

oxamic acid, oxalic acid, glyoxylic acid, glyoxal, CO2

FIGURE 18. Degradation of carbamazepine by ozonation [BQM: 1-(2-benzaldehyde)-4-hydro-(1H,3H)-quinazoline-2-one, BQD: 1-(2-benzalde-

hyde)-(1H,3H)-quinazoline-2,4-dione, BaQD: 1-(2-benzoic acid)-(1H,3H)-quinazoline-2,4-dione] (Andreozzi et al., 2002; McDowell et al., 2005).

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by ozonation at an applied ozone dose of 3 mg/L, pH 7.8and 23 �C. No further study has been reported on thedegradation of this anticonvulsant by ozonation or AOP.

Summary of Anticonvulsants and Anti-AnxietyAgents

Anticonvulsants and anti-anxiety agents reviewed hereshowed different reactivity toward ozonation and AOPs.This is primarily due to their structural diversity and theabsenceof commonreactive functional groups (seeFigure 16).Of the fourpharmaceuticals, carbamazepine is readilydegrad-able by either molecular ozone or hydroxyl radicals.Diazepamandprimidone are relatively resistant to ozonation,and the ozone treatment of buspirone has not been documen-ted. Carbamazepine degradation by ozonation and AOPs(H2O2/UV and TiO2/hn) has been extensively investigated,probably due to its relevance and persistence in the aqueousenvironment. Degradation pathways for this drug duringthese oxidative treatment processes have been proposed byseveral groups of investigators (Figures 18 and 19).While onereport has demonstrated the diminution of algal toxicity ofcarbamazepine by ozonation, the formation of acridine deri-vatives through advanced oxidation of the same drug (seeFigure 19) may warrant closer attention and possible cautionbecause of the potential ecotoxicity of these heteropolyaro-matic by-products. Further study may be required on the

other three drugs, as well as on those anticonvulsants/antide-pressants currently marketed and prescribed frequently, suchas paroxetine (Kwon and Armbrust, 2004).

ANTIPYRETICS AND NON-STEROIDAL ANTI-INFLAMMATORY DRUGS (NSAIDS)

Antipyretics and non-steroidal anti-inflammatorydrugs (NSAIDs) comprise one of the major classes ofpharmaceuticals commonly consumed in both prescrip-tion and non-prescription drugs. The NSAIDs covered inthis review include diclofenac, ibuprofen, indomethacin,naproxen, and salicylic acid. These are the drugs withanalgesic (reduce pain), antipyretic (reduce fever), andanti-inflammatory effects by inhibiting prostaglandinsynthesis by inhibition of cyclooxygenase (COX) (Merck& Co., 1999). Another popular antipyretic agent, para-cetamol, which is also known as acetaminophen andcovered in this section, has negligible anti-inflammatoryeffect and is not classified as NSAID. Molecular structureand molecular weight of these pharmaceuticals can befound in Figure 20. It should be noted that they are allacidic drugs.

Diclofenac. Diclofenac is a NSAID commonly usedfor musculoskeletal complaints such as arthritis. Thisacidic drug has been frequently detected in surface water

N

O NH2

carbamazepine

•OH/HO2•

N

O NH2

HO OO

N

O NH2

O

N

O

RR

N

O NH2

R = -COOH

N

O

OH

NH2

acridine

O

OH

OH

OH

OH

2-aminobenzoic acid salicylic acid catechol

organic acids

•OH •OH

•OH

N OHhydroxyacridine

FIGURE 19. Degradation of carbamazepine by H2O2/UV treatment (Vogna et al., 2004b). Compounds in square brackets were not

identified.

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(Ternes, 1998), groundwater (Sacher et al., 2001), andwastewater effluent (Ternes, 1998; Soulet et al., 2002;Ternes et al., 2003), with concentration levels up to 1.3mg/L (Ternes et al., 2003). A number of studies indicateda high reactivity of diclofenac toward ozonation (Zwienerand Frimmel, 2000; Ternes et al., 2002, 2003; Huber et al.,2005). For example, as low as 1 mg/L applied ozone wasenough to convert 1 mg/L diclofenac spiked in a floccu-lated surface water sample at pH 7.8 and 23 �C (Terneset al., 2002). Huber et al. (2003) reported a large secondorder rate constant (1 · 106 M�1�s�1) for the reaction ofozone with this pharmaceutical in the dissociated form atpH 5–10 and 20 �C. Another kinetic study indicatedsmaller rate constants ranging from 1.76 · 104 to 1.84 ·104 M�1�s�1 at pH 5–6 and 25 �C (Vogna et al., 2004a).The inconsistency may be due to the difference in meth-ods used by these two groups of investigators: the com-petition kinetics method employed by the formerresearchers and the direct measurement of residual drugusing HPLC by the latter. Various degradation intermedi-ates have been identified during the ozonation treatmentof 1 mM (296 mg/L) diclofenac at pH 7.0 (Vogna et al.,2004a). Hydroxylation of the aromatic rings and cleavageat the diphenyl amine nitrogen likely occur simulta-neously during the treatment (Figure 21), followed bythe ring opening and further mineralization. About 95%of chlorine from diclofenac molecules was released aschloride and more than 30% of initial TOC was removedafter 1.5 h of ozonation treatment (Vogna et al., 2004a).

Several AOPs have been also evaluated for the degra-dation of diclofenac. Zwiener and Frimmel (2000) noted apositive effect of hydrogen peroxide addition on thediclofenac conversion by ozone. A second-order rate con-stant for the reaction of hydroxyl radical and diclofenacwas determined as 7.2 · 109 M�1�s�1 (Huber et al., 2003).Vogna et al. (2004a) investigated H2O2/UV treatment ofthis pharmaceutical and compared the results with

ozonation and direct photolysis. Diclofenac (296 mg/L)was degraded by direct photolysis to some extent (>45%conversion) in 1.5 h using a 17-W low-pressure Hg lampemitting UV light at 254 nm with an intensity of 2.7 ·10–6 Einstein�s�1. The diclofenac degradation wasstrongly enhanced by the addition of hydrogen peroxide(170 mg/L; H2O2/UV AOP) to more than 90% conver-sion. About 50% of chlorine was recovered as chlorideion and about 40% of TOC was removed during thetreatment (Vogna et al., 2004a). Although the extent ofdiclofenac dechlorination was apparently less than thatachieved by ozonation, this is likely due to the incompletedegradation of parent compound and intermediates in theH2O2/UV treatment, and may be improved by optimizinghydrogen peroxide dose. A number of degradation inter-mediates were identified during the H2O2/UV treatmentof diclofenac as shown in Figure 21. Substitution of oneof two chlorine atoms with hydroxyl group was one of theunique degradation pathways of this UV assisted AOP(Vogna et al., 2004a). Andreozzi et al. (2004) reported theeffective detoxification of a mixture of pharmaceuticalscontaining 2.8 mg/L diclofenac and five other drugs byozonation and H2O2/UV treatment (see Ofloxacin, qui-nolone antibiotic for the details).

The Fenton-type process has an apparent disadvan-tage on the degradation of acidic drugs such as diclofenac(pKa=4.15) because the drug precipitates in an acidicmedium, which is required maintaining iron in solution.Packer et al. (2003) observed no conversion of diclofenacby Fenton treatment at pH 3.5 and 22 �C. While increas-ing pH can help the dissolution of the drug, precipitationof iron hydroxides occurs at high pH. To overcome thisdifficulty, it was suggested that the addition or generationof organic ligands or photo-reduction of colloidal iron tosoluble ferrous ion by UV irradiation might improve theperformance of Fenton-type oxidation of diclofenac(Perez-Estrada et al., 2005b). Ravina et al. (2002) investi-gated photo-Fenton oxidation of diclofenac using a con-centric photo-reactor with a 400 W low-pressure Hg lamp(l=254 nm). Majority of initial TOC (> 90%) could beremoved from 10 to 80 mg/L diclofenac solution within 1h by photo-Fenton treatment with 14 mg/L Fe2+ and 340mg/L H2O2 at pH 2.8 and 50 �C. No remark was given onthe solubility of diclofenac in this study, although it maybe possible that the drug solubility increased at the ele-vated temperature. It was noted that dark Fenton andH2O2 alone with UV irradiation (H2O2/UV) were muchless effective than the photo-Fenton treatment (Ravinaet al., 2002).

Perez-Estrada et al. (2005b) reported that neutral andacidic photo-Fenton treatment in buffered media was notvery effective for diclofenac degradation because of theprecipitation of either drug or iron hydroxides. Thephoto-Fenton treatment was performed in a pilot scalesolar photo reactor at 35 �C ([diclofenac]0=50 mg/L,concentration of H2O2 was maintained around 200 to

Cl

Cl

NHO

O

diclofenac (318.13)

O

OH

ibuprofen (206.28)

N

O

O

Cl

O OH

indomethacin (357.79)

OO

OH

naproxen (230.26)OH

O

OH

salicylic acid (138.12)

HO NHO

paracetamol (151.17)

Na+

FIGURE 20. Antipyretics and non-steroidal anti-inflammatory

drugs (NSAIDs).

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400 mg/L by continuous addition). A non-buffered sys-tem with an initial pH of 7.0 was superior to the bufferedsystems; however, some precipitation and subsequentre-dissolution of diclofenac was observed during the treat-ment. This was likely due to the pH drop (around pH 4.0)as a result of the generation of organic acids intermediates.Consequently, Perez-Estrada et al. (2005b) recommended apH control around 4.5 to 5 for the photo-Fenton treat-ment of diclofenac. A number of degradation intermedi-ates were detected during the photo-Fenton treatment(Ravina et al., 2002; Perez-Estrada et al., 2005a).Hydroxylation and subsequent quinoneimine formationdepicted in Figure 22 was considered as the major degra-dation pathway of diclofenac in the photo-Fenton process(Perez-Estrada et al., 2005a). These quinoneimine inter-mediates are likely decomposed into hydroquinone deriva-tives and aniline derivatives that have been detected inozonation and H2O2/UV treatment (Figure 21). Near-quantitative recovery of chlorine and nitrogen as chlorideand ammonium, respectively, was also confirmed duringthe treatment (Perez-Estrada et al., 2005a).

Titanium dioxide photocatalysis was also evaluated forthe degradation of diclofenac in the same solar photoreactor employed in the photo-Fenton process discussed

above (Perez-Estrada et al., 2005b). Complete conversionof 43 mg/L diclofenac was achieve with 0.2 g/L TiO2 in 200min, which was twice longer than the case of photo-Fentontreatment. Chloride and ammonium ions were detectedduring the TiO2/hn treatment (Perez-Estrada et al., 2005b).

Ibuprofen. Ibuprofen is a NSAID widely used for therelief of headache, rheumatoid arthritis, fever, andgeneral pain, and is an active ingredient of a number ofover-the-counter pain-relief drugs. Ibuprofen has beenfrequently detected in the aquatic environment (Halling-Sørensen et al., 1998; Ternes, 1998; Heberer, 2002; Kolpinet al., 2002; Boyd et al., 2003). Moderate reactivity of thisNSAID toward ozonation was reported (Zwiener andFrimmel, 2000; Ternes et al., 2003; Huber et al., 2005).A second-order rate constant for the molecular ozonereaction of ibuprofen was reported as 9.6 M�1�s�1(Huber et al., 2003), which is much lower than that ofdiclofenac. This is primarily due to the absence of reactivefunctional group or moiety such as amine and non-aro-matic carbon-carbon double bonds in the ibuprofenmolecule, besides the aromatic ring that is weakly acti-vated by the isobutyl group (see Figure 20).

Zwiener and Frimmel (2000) reported that ozonation(1 mg/L applied ozone) alone was not sufficient to

Cl

Cl

NHOH

O

diclofenac

Cl

Cl

NH2

O

OHHO

OHHO

Cl

Cl

O

OHHO

HOO

OHHO

OH

Cl

Cl

NHOH

OHOCl

Cl

NHOH

O

OHCl

Cl

NHOH

O

HO

+

O

OHH2N

Cl

Cl

NH2

ring opening

ring opening

(only by H2O2/UV)

OH

Cl

NHOH

O

(only by H2O2/UV)

FIGURE 21. Degradation of diclofenac by ozonation and H2O2/UV AOP (Vogna et al., 2004a).

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convert 2 mg/L ibuprofen in distilled water, resulting inonly 12% conversion in 10 min. However, increasing theozone dose and adding hydrogen peroxide (O3/H2O2

AOP) greatly improved the ibuprofen conversion. Theyalso investigated the effect of water matrix on the degra-dation of ibuprofen using a river water sample. The pre-sence of hydroxyl radical scavenger, bicarbonate ion andnatural DOC, in the river water significantly reduced theextent of ibuprofen conversion by O3/H2O2 AOP, imply-ing the importance of hydroxyl radical reactions. In theriver water, they noted that the treatment with 5.0 mg/Lapplied ozone and 1.8 mg/L hydrogen peroxide was suffi-cient to convert 2 mg/L ibuprofen in 10 min at pH 7.5 and10 �C (Zwiener and Frimmel, 2000). Two comparablesecond-order rate constants for the hydroxyl radical reac-tion of this pharmaceutical were reported: 6.5 · 109

M�1�s�1 at pH 3.5 and 22 �C by Fenton (Packer et al.,2003) and 7.4 · 109 M�1�s�1 at pH 7 and 25 �C by H2O2/UV (Huber et al., 2003). No further study was found onthe mineralization of ibuprofen by ozonation or AOP, orthe degradation pathway of this NSAID.

Indomethacin and Naproxen. Indomethacin andnaproxen are NSAIDs having methylated indole andnaphthalene rings, respectively. Like other acidicNSAIDs, the occurrence of these compounds in theaquatic environment have been reported (Halling-Sørensen et al., 1998; Ternes, 1998), although theirenvironmental concentrations are generally less thanthose of diclofenac and ibuprofen. Ternes et al. (2003)reported the effective abatement of 0.1 mg/L indometha-cin and 0.1 mg/L naproxen present in a biologicallytreated wastewater sample by ozonation at pH 7.2 withan applied ozone dose of 5 mg/L. Similar result wasreported by Huber et al. (2005) as well, although theconcentrations of these drugs were not specified. Asecond-order rate constant for the hydroxyl radical

reaction of naproxen has been reported as 9.6 · 109

M�1�s�1 at pH 3.5 and 22 �C using Fenton reaction togenerate hydroxyl radicals (Packer et al., 2003). Nofurther study has been found on the ozonation oradvanced oxidation treatment of these NSAIDs.

Salicylic Acid. Salicylic acid is a decomposition pro-duct of acetylsalicylic acid (aspirin), which is a commonover-the-counter NSAID. Salicylic acid itself, which isderived naturally from willow tree bark, has been alsoused as an antipyretic since ancient times, and still beenused as an additive of some skin-care products (Merck &Co., 1999). Only one study on salicylic acid removal byozonation has been reported. Khan et al. (2004) demon-strated partial degradation of salicylic acid by ozonationin the advanced water recycling demonstration plant inAustralia. The demonstration plant consists of lime clar-ification, dissolved air flotation, dual media filtration,ozonation, biological activated carbon filtration, micro-filtration, combined reverse osmosis/nanofiltration, andUV disinfection. About 60% of 0.65 mg/L salicylic acidfound in the influent of ozonation unit was degraded byozonation at an ozone dose of 17.5 mg/L for 15 min(Khan et al., 2004). No further study has been reportedon the degradation of salicylic acid or acetylsalicylic acidby ozonation or AOP.

Paracetamol. Paracetamol, also known as acetami-nophen, is a non-NSAID antipyretic agent used for therelief of fever, headaches, and other minor aches andpains. This drug is metabolized in the liver to mostlysulfate- and glucuronide-conjugates and excreted in theurine (Johnson and Plumb, 2005). Andreozzi et al.(2003a) showed that ozonation at pH 2 and 7 couldeffectively degrade 0.8 g/L paracetamol in 30 min withan ozone flow rate of about 72 g/h (reactorvolume=1.09 L). After two hours of ozonation, about20% and 30% of initial TOC was removed from the

Cl

Cl

NHOH

O

diclofenac

Cl

Cl

NHOH

O

OH

Cl

Cl

NOH

O

O

OH

Cl

NOH

O

O

Cl-

Cl

Cl

NOH

O

O

HO

Cl

Cl

N

OCl

Cl

N

O

OH

FIGURE 22. The major degradation pathways of diclofenac (via quinoneimine formation) by photo-Fenton process (Perez-Estrada et al., 2005).

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paracetamol solution at pH 2 and 7, respectively. A num-ber of degradation intermediates were formed during theozone treatment. The evolution of intermediates followstypical phenol ozonation pathways, such as hydroxyla-tion of phenol ring, anomalous ozonation to cleave aro-matic ring of hydroquinone, and decarboxylation byhydroxyl radicals [see Andreozzi et al. (2003a) for thedetails]. A simplified version of the proposed degradationscheme of paracetamol is shown in Figure 23. In addition,Andreozzi et al. (2003a) suggested the acid-catalyzedhydrolysis of amidic intermediates to occur. Second-order rate constants were determined for the reactionbetween molecular ozone and two forms of paracetamolas 1.41 · 103 M�1�s�1 (neutral) and 9.91 · 108 M�1�s�1(dissociated, pKa=9.36) (Andreozzi et al., 2003a).

Andreozzi et al. (2003a) also investigated the parace-tamol degradation by H2O2/UV AOP. They noted thatstrong enhancement of photo-degradation of this phar-maceutical by the addition of hydrogen peroxide. Morethan 90% of 1.51 mg/L paracetamol was converted in oneminute by the H2O2/UV treatment at pH 5.5([H2O2]0=170 mg/L, a low-pressure Hg lamp, l=254 nm).Mineralization monitored as TOC removal was alsoconfirmed during the treatment (Andreozzi et al.,2003a). A variety of degradation products were identifiedas shown in Figure 23 (Vogna et al., 2002; Andreozziet al., 2003a), and kinetic rate constants for paracetamoldegradation by hydroxyl radicals was determined as 2.2 ·109 M�1�s�1 at pH 5.5 (Andreozzi et al., 2003a). Besidesthe hydroxylation of the aromatic ring, the formation of

1,4-hydroquinone and 1,4-benzoquinone as a result ofeither para-hydroxylation (with respect to hydroxylgroup) or hydrogen abstruction from a phenoxyl OHcan occur during the H2O2/UV treatment of paracetamol(Figure 24) (Vogna et al., 2002).

In addition to ozonation and H2O2/UV AOP, anodicoxidation using a boron-doped diamond electrode hasbeen investigated for paracetamol degradation (Brillaset al., 2005). This electrochemical treatment employsadsorbed hydroxyl radicals generated at anode surfaceby electrolysis of water:

H2O! �OHads þHþ þ e�

OH� ! �OHads þ e�

The boron-doped diamond thin-film anode is much moreefficient in generating hydroxyl radicals than other con-ventional anodes such as platinum, PbO2, doped PbO2,doped SnO2 and IrO2, and enabled 70% to 98% miner-alization (monitored as TOC) of up to 948 mg/L para-cetamol after 4 h of treatment (Brillas et al., 2005). TheTOC removal by this electrochemical treatment was pH-independent in the pH range of 2.0 to 12.0. Oxalic acidand oxamic acid were detected as degradation intermedi-ates, which further degraded into carbon dioxide, ammo-nium, and nitrated (Brillas et al., 2005).

Summary of Antipyretics and NSAIDs

Antipyretics and NSAIDs reviewed here have shownrelatively low (e.g., ibuprofen) to high reactivity (e.g.,diclofenac) toward ozonation. Except for the cases ofdiclofenac and paracetamol, the information on thedegradation of these pharmaceuticals by ozonation andAOPs is generally limited. Degradation pathways havebeen proposed for diclofenac and paracetamol using var-ious treatment processes (Figures 21–24). Some kineticdata are available for diclofenac, ibuprofen, naproxen,and paracetamol. No potential ecological impact ofthese pharmaceuticals and their treatment by ozone andadvanced oxidation has been yet clearly addressed.Information is lacking on the treatment of other envir-onmentally relevant NSAIDs including ketoprofen(Ternes, 1998; Soulet et al., 2002) and mefenamic acid(Soulet et al., 2002), although their occurrences are lessfrequent than the ones covered in this review, such asdiclofenac and ibuprofen.

b-Blockers

Beta (b)-blockers are a class of drugs used to treat avariety of cardiovascular diseases, such as hypertension,coronary artery disease, and arrhythmias, by blocking theaction of epinephrine and norepinephrine on the b-adre-nargic receptors in the body, primarily in the heart(Merck & Co., 1999). The b-blockers covered in this

HO NHO

HO NHO

HOO3 or •OH

•OH

paracetamol

HO OH

O3

HO NHO

O OOHOH

HO OH

HO

•OHNH2

O NH2

O

•OH

OH

HO NHOO

OHO

CO2•OH•OH

HO

HOO

O

OH

HO

HOO

O

HO

HO OH

HO

OH OHO O

organic acids (e.g., oxalic acid, formic acid, glyoxalic acid, ketomalonic acid),

mineralization of organic acids

O3

O3 or •OH

O3 or •OHO3 or •OH

FIGURE 23. Degradation of paracetamol by ozonation and H2O2/

UV AOP (Vogna et al., 2002; Andreozzi et al., 2003).

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section include atenolol, celiprolol, metoprolol, propra-nolol, and sotalol. While propranolol and sotalol are non-selective b-blockers, the others are b1 antagonists (selec-tive b-blockers). Chemical structure and molecular weightof these pharmaceuticals are shown in Figure 25.

Atenolol, Celiprolol, Metoprolol, and Sotalol. Terneset al. (2003) demonstrated that ozonation at pH 7.2 waseffective for the degradation of five b-blockers, namelyatenolol, celiprolol, metoprolol, propranolol, and sotalol,present in an effluent from municipal sewage treatmentplant. Celiprolol, propranolol, and sotalol were appar-ently more reactive than atenolol and metoprolol: 5 mg/Lof applied ozone was sufficient to convert 0.28 mg/Lceliprolol, 0.18 mg/L propranolol, and 1.32 mg/L sotalol,whereas higher ozone dose (i.e., 10 to 15 mg/L) was

required to completely convert 0.36 mg/L atenolol and1.7 mg/L metoprolol (Ternes et al., 2003). No furtherstudy has been published on the degradation of theseb-blockers, except for propranolol (see below), by ozona-tion or AOP.

Propranolol. In addition to the study conducted byTernes et al. (2003) above, the high reactivity of propra-nolol toward ozonation was also reported elsewhere(Andreozzi et al., 2004). Complete conversion of 325.5mg/L propranolol was achieved by ozonation in 2 min atan absorbed ozone dose of as low as 4.75 mg/L, pH 7.4and 25 �C in a mixture of six pharmaceuticals, includingofloxacin, sulfamethoxazole, carbamazepine, clofibricacid, diclofenac, and propranolol. Besides ozonation,H2O2/UV treatment ([H2O2]0=5–10 mM, a low-pressureHg lamp, l=254 nm, 2.51 · 10�6 Einstein�s�1) was alsoeffective for the complete conversion of propranolol atpH 7.4 and 25 �C, whereas TiO2 photocatalysis was lesseffective (Andreozzi et al., 2004). Algal (Synechococcusleopoliensis) and protozoan (Brachionus calyciflorus) toxi-cities of the drug mixture were eliminated completely bythe ozone treatment. It was also shown that the H2O2/UVtreatment was also effective in algal toxicity reduction;however, it was less effective in protozoan toxicity reduc-tion. Finally, no toxicity reduction was achieved by theTiO2/hn treatment (Andreozzi et al., 2004). Degradationintermediates or by-products of propranolol were notdetermined.

Summary of b-Blockers

Despite their prevalence in the aquatic environment(Ternes, 1998; Sacher et al., 2001; Bendz et al., 2005),there have been very few reports on the degradation ofaqueous b-blockers by ozonation and AOPs. Neither

HO NHO

paracetamol

•OH

- H•

HO NHO

OH

HO OH

NH

O

HO NHO

OH

- H•

O NHO

•OH

O NO

O O

NH2

O

NH2

O

1,4-hydroquinone

1,4-benzoquinone

- H•

FIGURE 24. Formation of 1,4-hydroquinone and 1,4-benzoquinone from paracetamol by H2O2/UV AOP (Vogna et al., 2002).

O

NH2

ONH OH

atenolol (266.34)

O HN

O

N

OOH

NH

celiprolol (379.50)O

OOH

NH

metoprolol (267.37)

OOH

NH

propranolol (259.35)

HN

OH

NH

SOO

sotalol (272.36)

FIGURE 25. Beta-blockers.

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kinetics nor degradation by-product information is avail-able. Based on the limited information in the literature, itseems that all b-blockers reviewed here are fairly reactivetoward ozonation. These pharmaceuticals contain a sec-ondary amine group and a weakly/moderately activatedaromatic ring that are probable targets of molecularozone and hydroxyl radical attacks (see Figure 25). Ithas been reported that the toxicity of a drug mixturecontaining one of the b-blockers (propranolol) can beenremoved by ozonation and H2O2/UV AOP. However, thetoxicity of degradation by-products/intermediates of indi-vidual compounds is still unknown. Although the pre-sence and extent of ecological impact of environmentalb-blockers is uncertain, more study may be needed toconfirm no generation of potentially toxic intermediatesand by-products from the oxidative treatment of thesepharmaceuticals.

Cytostatic Drugs

Cytostatic drugs, also known as antineoplastic agents,are the pharmaceuticals used to treat various forms ofcancers. Some cytostatic drugs are also used to treatautoimmune diseases and to suppress transplant rejec-tions. There are a number of classes of cytostatic drugs,such alkylating agents, anti-metabolites, alkaloids, andanti-tumor antibiotics. The majority of these drugs inter-fere with mitosis (cell division) to selectively kill fast-growing tumor cells through a number of mechanismssuch as inhibition of DNA synthesis. Cytostatic drugsare cytotoxic by nature, as well as potentially carcino-genic and genotoxic (Merck & Co., 1999). The cytostaticdrugs reviewed here include six anthracyclines anti-tumorantibiotics (aclarubicin, daunorubicin, doxorubicin, epir-ubicin, idarubicin, and pirarubicin; Figure 26a), four anti-metabolites (5-fluorouracil, azathioprine, cytarabine, andmethotrexate; Figure 26b), and three nitrogen mustardalkylating agents (cyclophosphamide, ifosfamide, andmelphalan; Figure 26b, inset).

Anthracyclines (Aclarubicin, Daunorubicin, Doxo-rubicin, Epirubicin, Idarubicin, and Pirarubicin).Anthracyclines are chemotherapeutic antibiotics thatinhibit DNA and RNA synthesis by intercalating betweenbase-pairs of the DNA or RNA strands. They have acommon structure: 7,8,9,10-tetrahydro-5,12-naphthace-nedione moiety with a sugar chain (see Figure 26a).Castegnaro et al. (1997) evaluated Fenton reagent aswell as sodium hypochlorite and hydrogen peroxide toeffectively decompose six anthracyclines, including aclar-ubicin, daunorubicin, doxorubicin, epirubicin, idarubicin,and pirarubicin, that may be found in hospital wastes athigh concentrations. All anthracyclines (0.4 to 5 g/L; seeAppendix for the concentrations of individual drugs)were completely degraded after 1 h of treatment witheither 2.6% sodium hypochlorite or Fenton reagent(10.3 g/L Fe2+, 150 g/L H2O2, pH unknown), and therewas no residual mutagenicity, assayed by the Ames test

with or without metabolic activation. It was also notedthat hydrogen peroxide alone (150 g/L) was not veryeffective. No degradation products were detected byHPLC with a spectrofluorimeter, indicating effectivedestruction of the naphthacenedine moiety (Castegnaroet al., 1997). No further study has been reported on thedegradation of anthracyclines by ozonation or AOP.

Anti-Metabolites (Azathioprine, Cytarabine, 5-Fluor-ouracil, and Methotrexate). Anti-metabolites are cyto-static drugs that mimic purine or pyrimidine and keepthem from being incorporated into DNA or RNA.Azathioprine is an immunosuppressive agent widelyused in transplantations to control rejection reactions(Merck & Co., 1999). Rey et al. (1999) reported effectivedegradation of four anti-metabolite cytostatic drugs (269to 499 mg/L), including azathioprine, cytarabine, 5-fluor-ouracil, and methotrexate, by ozonation at pH 3 and 7(not azathioprine because of its poor solubility at neutralpH) within 1 h. They noted that there was no apparentdifference in the reaction rates at two pHs, indicating themajor contribution of molecular ozone reactions overhydroxyl radicals, and that the presence of multiplereactive sites, which is apparent from their molecularstructures (see Figure 26b). Mutagenicity of these anti-metabolites, with or without metabolic activation, werediminished after the ozone treatment (Rey et al., 1999).Degradation by-products were not identified in thisstudy.

Alkylating Agents (Cyclophosphamide, Ifosfamid,and Melphalan). As the name suggests, alkylatingagents are cytostatic drugs capable to covalently modifyelectronegative groups of DNA (i.e., DNA alkylation),and thus interfere DNA replication in tumor cells. Similarto the cases of anthracyclines above, Fenton oxidation, aswell as chlorination, of three alkylating agents, includingcyclophosphamide, ifosfamid, and melphalan, were inves-tigated as an effective treatment of hospital wastes con-taminated with these cytotoxic pharmaceuticals (Hanselet al., 1997). It was demonstrated that complete conver-sion of these alkylating agents could be achieved by thetreatment with Fenton reagent (10.3 g/L Fe2+, 150 g/LH2O2, pH unknown) within one hour. However, the for-mation of unidentified direct mutagens was noticed bythe Ames test in the treated cyclophosphamide formula-tion that had also contained 5% dextrose (Hansel et al.,1997). As no residual mutagenicity was found in the sameformulation without dextrose, this additive was complet-ing for the oxidant (hydroxyl radicals) with cyclopho-sphamide, resulting in incomplete destruction ofmutagenic degradation by-products.

Summary of Cytostatic Drugs

A large number of cytostatic drugs have been exam-ined for their degradation by ozonation or Fenton pro-cess in a very limited number of studies. Theconcentrations of the drugs treated therein were very

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high (269 mg/L to 27 g/L) as compared with their envir-onmentally relevant concentrations (<0.1 mg/L) (Halling-Sørensen et al., 1998; Ternes, 1998). This is due to the factthat these studies have been conducted to explore effec-tive treatment methods for highly polluted wastewaters.The evaluated treatment methods were generally effec-tive to degrade the cytostatic drugs and to remove muta-genicity, except for one case in which acyclophosphamide formulation was treated by theFenton process in the presence of 5% dextrose. Thisemphasizes the importance of monitoring the evolutionof degradation intermediates and ensuring the completedestruction of by-products during the treatment, espe-cially when other wastewater constituents that may com-pete for oxidants are present.

No kinetics data are available for the degradation ofcytostatic drugs by ozonation or AOP. Although theirenvironmental occurrence is rather rare because of theirlimited usage in the hospitals, effective treatment meth-ods should be investigated for the hospital wastes con-taining parent cytostatic drugs as well as their active(i.e., cytotoxic) metabolites excreted in the urine and

feces of patients as suggested by Castegnaro et al.(1997). In addition to the cytostatic drugs reviewedhere, further study is needed on the ozone and advancedoxidation treatment of many other drugs used for che-motherapy. One of the cytostatic drugs whose occur-rence has been described in the literature but notcovered here is bleomycin, a chemotherapeutic antibiotic(Halling-Sørensen et al., 1998).

Histamine H2-Receptor Antagonists

Histamine H2-receptor antagonists (H2 antagonist forshort) are the drugs used to block the action of histamineon the production of acid in the stomach. Two H2

antagonists have been detected in surface water samplesin the U.S., including cimetidine and ranitidine (Figure 27),at concentrations up to 0.58 mg/L and 0.01 mg/L, respec-tively (Kolpin et al., 2002).

Cimetidine. Latch et al. (2003) investigated the reac-tion of cimetidine (and ranitidine) with hydroxyl radicalsas one of the environmental degradation process. A sec-ond-order rate constant for protonated form of cimeti-dine (pKa=7.1; Figure 28) was determined as 6.5 · 109

M�1�s�1 using a competitive kinetics method with aceto-phenone as a reference compound (Latch et al., 2003).Several degradation products of cimetidine by Fentonoxidation were reported elsewhere as shown in Figure 29(Zbaida et al., 1986).

Ranitidine. A very high second-order rate constantfor the hydroxyl radical reactions was reported for theprotonated form of ranitidine (pKa=8.2; Figure 28) as1.5 · 1010 M�1�s�1 (Latch et al., 2003). Addamo et al.(2005) reported the degradation of this H2-antagonist byTiO2 photocatalysis. Almost complete conversion of50 mg/L ranitidine was achieved by the TiO2/hn treat-ment in 1 h with 0.4 g/L of catalyst and a 125 W mediumpressure Hg lamp with a photon flux of 8.5 mW�cm�2 at40 �C (Addamo et al., 2005). In addition, about 60% ofTOC was removed by prolonged treatment for 5 h. Onthe other hand, direct photolysis was much less effectivein ranitidine conversion and mineralization (Addamoet al., 2005). Hydroxylation of furan ring and S-oxidationwere suggested as possible degradation pathways,although they were not experimentally demonstrated.

Summary of Histamine H2-Receptor Antagonists

Very few studies have been published on the degrada-tion of histamine H2-receptor antagonists by AOPs, andnone has been reported on their ozone treatment. Sincetwo H2 antagonists appear to be fairly reactive towardozonation because of the presence of a few reactive func-tional groups/moieties such as furan ring, amine, andthioether sulfur (see Figure 27), ozonation may be agood treatment process for surface water contaminatedwith these pharmaceuticals. Further study is required toaddress the potential ecotoxicity of environmental H2

O

OOH OH O

OH

O O

ON

O

O

OHO

OO aclarubicin (811.88)

O

OO OH O

OH

ONH2OH

daunorubicin (527.53)

OOH

O

OO OH O

OH

ONH2OH

OOH OH

doxorubicin (543.52)

O

OO OH O

OH

ONH2

OOH OH

HO

epirubicin (543.52)

O

O OH O

OH

ONH2OH

OOH

idarubicin (497.50)

O

OO OH O

OH

ONH2O

OOH OH

O

pirarubicin (627.64)

FIGURE 26a. Cytostatic drugs, anthracyclines.

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antagonists as well as their degradation by-products aris-ing from the oxidative treatment such as ozonation andadvanced oxidation.

HORMONES AND ORAL CONTRACEPTIVES

A number of natural and synthetic steroid hormonesare used in treatment of various types of medical condi-tions such as menopausal symptoms, growth hormonedeficiency, hypothyroidism, and some forms of cancers

(Arcand-Hoy et al., 1998). Some natural and syntheticestrogen hormones are also used as oral contraceptives.Four types of estrogen hormones are covered in thisreview, including 17b-estradiol, estrone, 17a-ethinylestra-diol, and diethylstilbestrol (Figure 30). While the formertwo are natural (endogenous) estrogens, the latter two aresynthetic estrogens.

17b-Estradiol. 17b-Estradiol is an endogenous estro-gen responsible for the development of female secondarysex characteristics and reproduction. In addition to itsendogenous occurrence, this natural estrogen is

N

N S

N

N N

HN

NO2

azathioprine (277.26)

N

N

NH2

O

O

HO

OH

HO

cytarabine (243.21)

NH

HNF

O

O

5-fluorouracil (130.08)

N

N

N

NH2N

NH2

N

O

HN

O OH

O

OH

methotrexate (454.44)

PO

NHN

O

ClCl

cyclophosphamide (261.09)

PO

NNH

O

Cl Cl

ifosfamid (261.09)

NCl

Cl

OHNH2

O

melphalan (305.20)

FIGURE 26b. Cytostatic agents, anti-metabolites and nitrogen mustard alkylating agents (inset).

NHN S

HN NH

NN

cimetidine (252.34)

OS

HN

NHN

NO2

ranitidine (314.40)

FIGURE 27. Histamine H2-receptor antagonists.

NHHN S

HN NH

NN

OS

HN

NHHN

NO2

FIGURE 28. Protonated forms of cimetidine (left) and ranitidine

(right).

NHN S

HN NH2

NN

N-desmethylcimetidine

NHN S

HN NH

N

O

NH2

cimetidine guanylurea derivative

NHN S

HN NH

NN

O

cimetidine sulfoxide

NHN S

HN NH

NN

OH

O

5-hydroxylmethylimidazole cimetidine sulfoxide

FIGURE 29. Degradation products of cimetidine by Fenton pro-

cess (Zbaida et al., 1986).

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manufactured and used in oral contraceptives and hor-mone replacement therapy in large quantities. 17b-Estradiol is metabolized in the body to a number ofmetabolites, including estrone, hydroxylated estrones, 2-methoxyestrone, estriol, and their conjugates, and subse-quently excreted in the urine (Arcand-Hoy et al., 1998).Deconjugation of sulfate and glucuronide derivatives ofestradiol in sewage treatment plant is also suggested(Arcand-Hoy et al., 1998; Belfroid et al., 1999). 17b-Estradiol has been frequently detected in the aquaticenvironment (Ternes, 1998; Kolpin et al., 2002; Snyderet al., 2003), and is considered as a major contributor ofestrogenic activity found in municipal sewage treatmentplant effluent (Onda et al., 2002).

Several groups of investigators reported the effectivedecomposition of 17b-estradiol by ozonation (Ondaet al., 2002; Alum et al., 2004; Kim et al., 2004; Huberet al., 2005). For example, ozonation at 2 mg/L appliedozone dose was sufficient to convert 0.5 mg/L 17b-estradiolspiked in biologically treated municipal wastewater at pH7 and 16 �C (Huber et al., 2005). Some studies alsoaddressed the effects of ozone treatment on estrogenicactivity, which is assayed by a number of methods suchas recombinant yeast assay (Onda et al., 2002), estrogencompetition binding assay (Kim et al., 2004), E-screenassay (Alum et al., 2004), MCF-7 cell proliferation assay(Liu et al., 2005), and ER-binding assay combined withultrafiltration (Liu et al., 2005). Generally, all of thesestudies indicated that estrogenic activity of 17b-estradiolsolution decreased upon the ozone treatment. A slightincrease of estrogenicity assayed by E-screen was notedin the first five minutes of ozone treatment (Alum et al.,2004), indicating the formation of more potent estrogenicintermediates, although the estrogenicity decreased duringthe subsequent 5 min of treatment. Huber et al. (2004)identified four degradation intermediates of 17b-estradiolby ozonation in the presence of a hydroxyl radical

scavenger (Figure 31). It is apparent that ozone moleculesattack and cleave hydroxylated aromatic ring of this estro-gen to produce these intermediates. Huber et al. (2004)suggested that the destruction of aromatic ring virtuallydiminished estrogenicity of estrogens including 17b-estradiol.

It was also documented the effectiveness of severalAOPs including O3/H2O2 (Shishida et al., 2000; Ondaet al., 2002) H2O2/UV (Rosenfeldt and Linden, 2004)and TiO2/hn processes (Ohko et al., 2002) on the degra-dation of 17b-estradiol and the removal of associatedestrogenicity. Shishida et al. (2000) showed that in addi-tion to estrogenicity, cytotoxicity, mutagenicity, and gen-otoxicity initially present in a secondary effluent fromdomestic wastewater treatment plant containing 17b-estradiol (concentration not shown) was reduced by theO3/H2O2 treatment with 30 mg/L ozone and 2 mg/LH2O2. A second-order rate constant for the reaction of17b-estradiol and hydroxyl radical was determined as1.41 · 1010 M�1�s�1 using a competitive kinetics methodwith isopropyl alcohol as a reference compound(Rosenfeldt and Linden, 2004). Ohko et al. (2002) demon-strated complete mineralization of 0.272 mg/L 17b-estradiol by TiO2 photocatalysis with 1.0 g/L TiO2 anda 200-W Hg-Xe lamp (with a 365 nm band-pass filter) in 3h. They also identified several degradation intermediatesdetected during the TiO2 photocatalytic treatment of 17b-estradiol, including 10�-17b-dihydroxy-1,4-estradien-3-one, androsta-4,16-dien-3-one, and testosterone (Figure 32),although it is unlikely to occur the introduction of methylgroup on the C10 position by the TiO2/hn treatment. Ohko etal. (2002) suggested that hydroxyl radical attacked the hydro-xylated aromatic ring of 17b-estradiol to initiate its degrada-tion. In addition to the three AOPs discussed here, thedegradability of 17b-estradiol by photo-Fenton-like processwas briefly compared with that of three estrogens, includingdiethylstilbestrol, 17a-ethinylestradiol, and estrone (Feng

HO

17β-estradiol (272.38)

HO

O

estrone (270.37)

HO

OHOH

17α-ethinylestradiol (296.41)

HO

OH

diethylstilbestrol (268.35)

FIGURE 30. Hormones and oral contraceptives.

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et al., 2005). The order of degradability was diethylstilbestrol> 17b-estradiol > 17a-ethinylestradiol > estrone([estrogen]0=5 mg/L, 0.5 mg/L Fe3+, 28.3 mg/L H2O2, a250-W metal halide lamp, pH 3.0).

Estrone. Estrone is one of the natural metabolites of17b-estradiol having weaker estrogenic activity than theparent compound (Arcand-Hoy et al., 1998). Sulfate esterof this natural estrogen is one of the active ingredients ofconjugated equine estrogens used for the treatment ofhormone replacement therapy. Estrone sulfate undergoeshydrolysis to allow its absorption by the gastrointestinaltract. Estrone has been detected in treated domestic was-tewater and surface water nearly as frequently as 17b-estradiol (Belfroid et al., 1999; Kolpin et al., 2002).

Degradation of aqueous estrone by ozonation andphoto-Fenton-like treatment was investigated. Completeconversion of 0.015 mg/L estrone present in a municipalsewage treatment plant effluent was achieved by ozona-tion at an applied ozone dose of 5 mg/L and pH 7.2

(Ternes et al., 2003). Huber et al. (2004) proposed adegradation pathway identical to that of 17b-estradiolby ozonation (Figure 31). Huber et al. (2005) also showedthat estrone was the least reactive toward ozonationamong three estrogens tested, including 17b-estradiol,17a-ethinylestradiol, and estrone. Consistent, lower reac-tivity of estrone was observed in photo-Fenton-like treat-ment as well (Feng et al., 2005), although the reason forthis relative recalcitrance was not given in either case.Complete conversion and about 15% mineralization(monitored as CO2 evolution) of 5 mg/L estrone wasachieved in 160 min by the photo-Fenton-like treatmentwith 1.1 mg/L Fe3+, 56.6 mg/L H2O2, and a 250-W metalhalide lamp at pH 3.0. Six unidentified degradation inter-mediates, which are more polar than the parent com-pound, were detected by HPLC with a UV detectorduring the treatment (Feng et al., 2005).

17a-Ethinylestradiol. 17a-Ethinylestradiol is a syn-thetic estrogen commonly used in oral contraceptives.Introduction of ethinyl group to estradiol inhibit its meta-bolism in the liver, resulting in enhanced bioavailabilityand effectiveness (Arcand-Hoy et al., 1998). At the sametime, this synthetic estrogen is known to be more resistantthan 17b-estradiol against biodegradation (Ternes et al.,1999; Jurgens et al., 2002). Their frequent occurrences insewage treatment plant effluents and surface water havebeen also reported, however, at concentrations generallylower than those of natural estrogens, such as estrone and17b-estradiol (Belfroid et al., 1999; Ternes et al., 1999).

Huber et al. (2003) demonstrated very rapid reactionof 17a-ethinylestradiol and molecular ozone with a max-imum second-order rate constant of 7 · 109 M�1�s�1 at20�C at pH 10. The reaction rate is pH dependent; thedissociated (phenolate) form of this compound(pKa=10.4) is more reactive than the neutral moleculetoward molecular ozone. Actually, this synthetic estrogenhas two reactive moieties with different reactivity: a

HO

17β-estradiol

OH

HO

O

estrone

O

HO

O

OHO

O

O

OH

HO

O

OHOH

HO

O

HO

FIGURE 31. Proposed degradation pathways for 17b-estradiol and estrone by ozonation (Huber et al., 2004).

O

17β-estradiol

OH

HO

O O

OH

HO

OH

12

3 4 5 678

10911

12 13

14 15

1617

18

1 2 3

FIGURE 32. Proposed degradation intermediates of 17b-estra-

diol by TiO2 photocatalysis (Ohko et al., 2002): 1) 10�-17b-dihy-droxy-1,4-estradien-3-one, 2) androsta-4,16-dien-3-one, and 3)

testosterone.

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hydroxyl aromatic ring (kO3=3 · 106 M�1�s�1 at pH 7)and an ethinyl group (kO3=200 M�1�s�1) (Huber et al.,2004). These moieties are attacked by molecular ozonesimultaneously, resulting in a complex mixture of degra-dation products shown in Figure 33. A negative effect ofdissolved organic matter in natural water samples on 17a-ethinylestradiol conversion was reported (Huber et al.,2003). Similar to the case of 17b-estradiol, decrease ofestrogenic activity upon ozonation of this synthetic estro-gen was also demonstrated (Alum et al., 2004; Huberet al., 2004; Liu et al., 2005).

Rosenfeldt and Linden (2004) observed the improve-ment of UV photodegradation of 17a-ethinylestradiol(concentration not specified) by the addition of hydrogenperoxide (i.e., H2O2/UV AOP). Two consistent second-order rate constants were reported for the hydroxyl radi-cal reaction of 17a-ethinylestradiol: 9.8 · 109 M�1�s�1(Huber et al., 2003) and 1.08 · 1010 M�1�s�1(Rosenfeldt and Linden, 2004), both using a competitionkinetics method. A kinetic model was developed to pre-dict the degradation of this synthetic estrogen by H2O2/UV AOP under various reaction conditions, such aswater quality, concentration of hydrogen peroxide, andtype of UV sources (Rosenfeldt and Linden, 2004). Otherthan H2O2/UV, photo-Fenton-like process has been eval-uated for the degradation of 17a-ethinylestradiol (Fenget al., 2005) as described here (see 17b-Estradiol).

Diethylstilbestrol. Relatively high degradability ofdiethylstilbestrol, a synthetic estrogen, by photo-Fenton-like

process was briefly described (see 17b-Estradiol) (Feng et al.,2005). No further study was found on the degradation of thissynthetic estrogen by AOP or ozonation. It should be notedthat although diethylstilbestrol once had been often pre-scribed to prevent miscarriages, the use of this highly potentestrogen is now limited to the treatment of prostate cancerbecause of its carcinogenicity (Arcand-Hoy et al., 1998).

Summary of Hormones and Contraceptives

Two estrogenic steroid hormones and two syntheticestrogens have been studied for their degradation by ozo-nation and several AOPs. It can be concluded that thesenatural hormones and synthetic drugs are generally reactivetoward the oxidative degradation, although some studieshave suggested that estrone is relatively resistant.Estrogenicity associated with these compounds can bereduced by ozonation and O3/H2O2 (most likely by otherAOPs as well) to virtually undetectable levels. The degrada-tion of these estrogens is initiated by the hydroxylation ofhighly reactive aromatic ring followed by the ring openingand further degradation (Figures 31 and 33). Since thephenolic A ring (i.e., the hydroxylated aromatic ring) ofthese compounds is the structural component responsiblefor the high-affinity binding to the estrogen receptor(Arcand-Hoy et al., 1998), its modification and cleavagelikely lead the diminution of estrogenicity of degradationproducts produced by ozonation or AOPs. In addition tothe four estrogenic compounds, there are other steroidhormones for therapeutic use such as equilin, a horse

O

HO

O

HO O

HO

O

HO O

HO OH

OHO

O

HO O

O OH

O

O

O

HO

O

HOHO

O

HO

OH

O

HO

HO

O

HO

OH

O

O

O

HO

O

HO O

HO

O

O

O

HO

O

O

O

FIGURE 33. Some ozonation by-products of 17a-ethinylestradiol. The compounds in brackets were not detected, but likely formed during

ozonation, based on the results of model compound degradation [see Huber et al. (2004) for the details of the degradation mechanisms].

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estrogen often used in combination with estrone for hor-mone replacement therapy, and testosterone have beendetected in surface water (Kolpin et al., 2002), althoughtheir degradation by ozonation or AOP has yet beenstudied.

Lipid Regulators

Fibrate lipid regulators are the pharmaceuticals usedfor a range of metabolic disorders, mainly hypercho-lesterolemia (Merck & Co., 1999). They are phenox-yalkanoic acid derivatives, either free acid or esters,and accelerate the clearance of very-low-density lipo-proteins (VLDL). Bezafibrate, clofibrate, fenofibrate,and gemfibrozil, and their hydrolyzed metabolites,including clofibric acid and fenofibric acid (Figure 34)have been frequently found in the aquatic environmentin a number of countries (Buser et al., 1998; Ternes,1998; Kolpin et al., 2002; Soulet et al., 2002; Bendzet al., 2005). It should be noted that bezafibrate andgemfibrozil are acidic drugs on their own, while clofi-brate and fenofibrate become acidic metabolites afterthe hydrolytic biotransformation.

Bezafibrate. It was shown that bezafibrate was rela-tively resistant to the degradation by ozone as comparedwith other pharmaceuticals such as carbamazepine, diclo-fenac, and 17a-ethinylestradiol (Ternes et al., 2002;Huber et al., 2003). A second-order rate constant for thereaction of bezafibrate and molecular ozone was deter-mined as 590 M�1�s�1 at pH 5 to 10 and 20 �C (Huberet al., 2003). The negative impact of DOC in naturalwater on bezafibrate conversion by ozonation wasobserved. Huber et al. (2005) also suggested the impor-tance of hydroxyl radical reactions (k�OH=7.4 · 109

M�1�s�1 at pH 7 and 25 �C) during the ozonation ofthis pharmaceutical. No degradation by-product wasidentified in any of these studies.

Clofibric Acid. Clofibric acid, a metabolite of fibratelipid regulators clofibrate and etofibrate, was detected indomestic wastewater effluent nearly 30 years ago (Higniteand Azarnoff, 1977). More recently, the persistence andmobility of this acidic drug metabolite in the aquaticenvironment has been recognized (Buser et al., 1998).This compound is relatively resistant to ozone treatment.For example, only 8% of 2 mg/L clofibric acid was con-verted by ozonation in 10 min at an applied ozone dose of1 mg/L and pH 7, by which more than 96% of 2 mg/Ldiclofenac, a NSAID, was converted (Zwiener andFrimmel, 2000). Similar results were also reported else-where (Ternes et al., 2002, 2003; Andreozzi et al., 2004;Huber et al., 2005). Huber et al. (2005) reported a rela-tively small second-order rate constant of <20 M�1�s�1for the molecular ozone reaction of this acidic drug meta-bolite. It can be suggested that weak deactivation byp-chloro group (Beltran, 2003) (see Figure 34) is respon-sible to this low reactivity of aromatic ring towardozonation.

Degradation of clofibric acid by ozone can beenhanced by increasing ozone dose (Ternes et al., 2003),adding hydrogen peroxide (i.e., O3/H2O2 AOP) (Zwienerand Frimmel, 2000), or elevating pH (Andreozzi et al.,2003b). Latter two approaches indicate the importance ofhydroxyl radical reactions for the degradation of thisacidic drug metabolite. The apparent second-order rateconstants for the clofibric acid decomposition were pHdependent; 29.8 M�1�s�1 at pH 2.0 and 2550 M�1�s�1 atpH 6.5, both in the absence of hydroxyl radical scavenger(Andreozzi et al., 2003b). Andreozzi et al. (2003b) alsodemonstrated that complete conversion and 49% miner-alization, monitored as TOC removal, of 322 mg/L clo-fibric acid could be achieved by continuous ozonation(0.48 mg O3/L in bulk liquid) in one hour at pH 5.0 and25 �C. Almost complete dechlorination (>90%) was alsoobserved under the same conditions.

ClHN

O

O

O

OH

bezafibrate (361.82)

O

Cl

OH

O

clofibric acid (214.65)

O

Cl

O

O

clofibrate (242.70)

O

ClO

O

O

fenofibrate (360.84)

O

ClO

O

HO

fenofibric acid (318.75)

O

O

OH

gemfibrozil (250.34)

FIGURE 34. Fibrate lipid regulators and metabolites.

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In addition to ozonation and O3/H2O2 process, H2O2/UV and TiO2 photocatalysis have been evaluated for thedegradation of clofibric acid. Andreozzi et al. (2003b)demonstrated the more than 90% conversion of 322 mg/L clofibric acid by H2O2/UV treatment in 1 h with 34 g/LH2O2 and a 17 W low-pressure lamp (l=254 nm, 2.7 ·10�6 Einstein�s�1) at pH 5.0 and 25 �C. More than 80%of chlorine was recovered as chloride during the treat-ment, whereas TOC removal was modest. A second-orderrate constant for the hydroxyl radical reaction of clofibricacid was determined as 2.38 · 109 M�1�s�1, which isindependent of pH in the range of 4 to 7 (Andreozziet al., 2003b). This rate constant is smaller than the oneestimated in a separate study (4.7 · 109 M�1�s�1) usingFenton reaction to generate hydroxyl radicals at pH 3.5and 22 �C (Packer et al., 2003). A kinetic model wassuccessfully developed to predict the clofibric acid degra-dation at more dilute acid solution (10.7 mg/L) that repre-sents a sewage treatment plant effluent. Andreozzi et al.(2004) also investigated the detoxification of a drug mix-ture that contained clofibric acid by ozonation, H2O2/UV, TiO2 photocatalysis (see Propranolol, b-blocker forthe results).

A series of studies was published on the clofibric aciddegradation by TiO2 photocatalysis (Doll and Frimmel,2004, 2005a, 2005b, 2005c). More than 90% of 0.53 mg/Lclofibric acid was converted in five minutes with 80 mg/LTiO2 (P25) and a 1000-W xenon short-arc lamp as asource of simulated solar UV rays (1.35 · 104

Einstein�m�2�s�1, l <400 nm) at pH 6.5 (Doll andFrimmel, 2004). The degradation of clofibric acid wasdependent on the initial concentration of substrate;degradation efficiency was higher as the initial clofibricacid concentration decreased, indicating that the capacityof catalyst is a limiting factor (Doll and Frimmel, 2004).

Degradation efficiency was also decreased in the presenceof NOM, which completes with clofibric acid for theactive site of the catalyst (Doll and Frimmel, 2005a). Itwas also shown that clofibric acid was more readilydegradable by TiO2/hn than two X-ray contrast media,namely iomeprol and iopromide, but less degradable thanan anticonvulsant carbamazepine. A successful pilot-scaleinvestigation of the continuous treatment of this drugmetabolite by the TiO2/hn process was reported using across-flow microfiltration to retain catalyst (Doll andFrimmel, 2005b). A number of degradation productswere identified and degradation pathway was proposedas shown in Figure 35 (Doll and Frimmel, 2004). Amongthe degradation by-products, 4-chlorophenol was moreresistant to further degradation and accumulated in thesolution during the TiO2/hn treatment (Doll andFrimmel, 2004), which may raise concerns over the resi-dual toxicity, although it was not addressed therein.

Fenofibric Acid. Complete elimination of 0.13 mg/Lfenofibric acid, a metabolite of lipid regulator fenofibrate,in a sewage treatment plant effluent was achieved byozonation with an applied ozone dose of 10 or 15 mg/Lat pH 7.2 (Ternes et al., 2003). No further study wasfound on the degradation of fenofibric acid or its parentcompound fenofibrate by ozonation or AOP.

Gemfibrozil. Huber et al. (2005) mentioned the pre-sence of another fibrate lipid regulator, gemfibrozil, in amunicipal wastewater effluent that was subsequently trea-ted by ozonation; however, no description was given onthe removal efficiency of this pharmaceutical. It can beanticipated that gemfibrozil is probably more reactivetoward ozonation than clofibric acid because of theabsence of chlorine on the aromatic ring, although aro-matic hydroxylation is somehow hindered by the twomethyl groups (see Figure 34).

O

Cl

OH

O

clofibric acid

OH

Cl+

OH

OHO

OH

O

OH

O

HO

4-chlorophenol

O

HO

OH

O

2-(4-hydroxyphenoxy)-isobutyric acid

OH

HO

hydroquinone

OH

Cl OH

4-chlorocatachol

ring opening and

mineralization

FIGURE 35. Degradation of clofibric acid by TiO2 photocatalysis (Doll and Frimmel, 2004).

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Summary of Lipid Regulators and Metabolites

It can be concluded that lipid regulators and metabo-lites, especially clofibric acid, are not very reactive towardozonation. On the other hand, advanced oxidation isapparently more suitable for the degradation of thesepharmaceuticals. Degradation of clofibric acid has beenextensively studied and various treatment processes havebeen evaluated. Whereas some kinetic data on the degra-dation of this drug are available (see Appendix), degrada-tion pathway is only proposed for TiO2 photocatalysis(Figure 35). Such information is largely missing for theother lipid regulators. Generation and accumulation oftoxic chlorinated intermediates such as 4-chlorophenolwas observed in one study on clofibric acid degradationby TiO2/hn process (Doll and Frimmel, 2004).

X-RAY CONTRAST MEDIA

X-ray contrast media are the non-therapeutic medicalagents used to enhance the visibility of internal body struc-tures by X-ray imaging technologies. Triiodinated benzenederivatives are widely used for this purpose because of theirhigh water solubility and biologically inert (i.e., stable)nature. However, these biologically stable compounds arealso proven highly resistant to biodegradation and thuspersistent in the aquatic environment (Ternes and Hirsch,2000; Drewes et al., 2001). In addition, the removal oftriiodinated X-ray contrast media by sorption is poor dueto their high hydrophilicity. These contrast agents areexcreted with no marked metabolic transformation andcontribute to the increase of absorbable organic halogens(AOX) in hospital wastewater (Gartiser et al., 1996), as wellas high concentration of absorbable organic iodine (AOI)in sewage treatment plant effluent (5 to 40 mg I/L) (Dreweset al., 2001), surface water and raw drinking water (11 to 13mg I/L) (Putschew et al., 2000). Although it has been sug-gested that no environmental risk is evident based on abattery of acute and chronic toxicity test data of triiodi-nated X-ray contrast media (Steger-Hartmann et al., 1999),their relatively high environmental concentrations, persis-tence, mobility, as well as the lack of information regardingpotential sub-lethal effects have generated interests in theapplication of advanced water and wastewater treatmentprocesses including ozonation and AOPs. The X-ray mediainvestigated include diatrizoate, iomeprol, iopamidol,iopentol, and iopromide (Figure 36). Diatrizoate is theonly ionic contract media, while the others are non-ionic.

Diatrizoate. Diatrizoate is an anionic triiodinatedX-ray contrast medium. Ternes et al. (2003) reportedthe high recalcitrance of this compound against ozona-tion. No conversion of 5.7 mg/L diatrizoate in a sewagetreatment plant effluent was observed by ozonation at anapplied ozone dose of 5 mg/L at pH 7.2. Increasing ozonedose to 15 mg/L, addition of hydrogen peroxide (10 mg/LH2O2 and 10 mg/L O3; O3/H2O2), or UV irradiation (15mg/L O3 and a 110-W low-pressure UV unit, 254 + 185

nm; O3/UV) slightly enhanced the degradation of dia-trizoate to 14%, 25%, and 36%, respectively. Huber etal. (2005) also noted the negligible and very little reactiv-ity of diatrizoate toward molecular ozone and hydroxylradicals, respectively. No further study was found on thedegradation of diatrizoate by ozonation or AOP. Itshould be noted that ionic contrast media like diatrizoateare less commonly used nowadays than non-ionic coun-terparts as relatively high nephrotoxicity of the formercompounds has been recognized (Soejima et al., 2003).

Iomeprol. Ternes et al. (2003) demonstrated partialdegradation of iomeprol (from 34% to 90%) by ozonationand two types of ozone-based AOPs (see Diatrizoate forreaction conditions). This non-ionic triiodinated X-ray con-trast medium was more degradable than diatrizoate. Acomparable result was also reported by Huber et al.(2005). Unlike the case of diatrizoate, UV irradiation didnot improve the conversion of iomeprol by ozonation(Ternes et al., 2003). Doll and Frimmel (2004) investigatedthe degradation of iomeprol by TiO2 photocatalysis. About68% of 4.1 mg/L iomeprol was converted and about 40%ofquantitative iodine was released as iodide after 3 min ofTiO2/hn treatment with 500 mg/L TiO2 (Hombikat UV100) and simulated solar radiation using a 1000-W Xeshort-arc lamp at pH 6.5 and 20 �C. As compared with theiomeprol conversion and deiodination, DOC removal wasvery slow, indicating the generation of organic by-productsand intermediates (Doll and Frimmel, 2004), although theiridentities were unknown. Similar to the cases of carbamaze-pine and clofibric acid, negative impacts of NOM and otherorganic constituents in water was observed during the TiO2/hn treatment of iomeprol (Doll and Frimmel, 2005a).

Iopamidol. Degradability of iopamidol by ozone-based processes was equivalent to that of iomeprol(Ternes et al., 2003; Huber et al., 2005). No furtherstudy was found on the degradation of this non-ionicX-ray contrast medium by ozonation or AOP.

Iopentol. Sprehe et al. (2001) demonstrated completeAOX removal and partial mineralization (up to 80% asTOC removal) of 785 mg/L iopentol by H2O2/UV AOPin 3 and 5.5 h, respectively, with a 1 kW high-pressure Hglamp (output adjusted to 500 W) at pH 6.6. Hydrogenperoxide was supplied based on the COD of the solution.It was suggested that deiodination mostly occurred byUV photolysis of the parent compound:

Iopentolþ hn at 242 nmð Þ ! deiodinated intermediatesþ 3I�

I� þH2O orsubstrateð Þ ! I� þ �OHþHþ

Subsequently, two iodides react in the presence of amolecule of hydrogen peroxide and two protons,forming elemental iodine, which can be stripped outby air:

2I� þH2O2 þ 2Hþ Ð I2 " þ2H2O

386 K. Ikehata et al. December 2006

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Because of the high UV absorbance of iodine in theaqueous solution, iopentol mineralization was inhibited,although it was found that air stripping could solve thisproblem (Sprehe et al., 2001). They also suggested thatthe iodine recovery from the exhaust air by the use ofsolvent or sublimation might be feasible as about 85% ofelemental iodine could be recovered this way.

Iopromide. Iopromide is another non-ionic triiodi-nated X-ray contrast medium resistant to oxidative degra-dation. However, as compared with other triiodinatedcontrast media, this compound was apparently more reac-tive, but very slightly, toward ozonation (Ternes et al.,2003; Huber et al., 2005). As the second-order rate con-stant for the reaction between iopromide and molecularozone is very small: less than 0.8 M�1�s�1 at pH 5 to 10and 20 �C, the degradation of this compound mostly occurby hydroxyl radicals, although the rate constant for thehydroxyl radical reaction is also fairly small: 3.3 · 109

M�1�s�1 at pH 7 and 25 �C (Huber et al., 2003).Like the case of iopentol, effective AOX removal and

mineralization of iopromide by H2O2/UV AOP wasdemonstrated (Sprehe et al., 2001). Major contributionof direct UV photolysis to the deiodination was con-firmed in a separate photolysis experiment withouthydrogen peroxide addition. In addition to ozone-basedand H2O2/UV AOPs, TiO2/hn process was evaluated forthe iopromide degradation. It was shown that the degra-dation rate of this contrast medium was comparable tothat of iomeprol (Doll and Frimmel, 2004). Degradation

by-products or intermediates were not identified in any ofthese studies.

Summary of X-ray Contrast Media

It is evident that triiodinated X-ray contrast media arethe most refractory class of pharmaceutical/medicalagents reviewed in this review. Anionic diatrizoate isparticularly resistant to ozone-based treatment, whereasnon-ionic contrast media (iomeprol, iopamidol, iopentol,and iopromide) are somewhat degradable, especially byAOPs such as H2O2/UV and TiO2/hn. It has been shownthat deiodination of these contrast media occurs chieflyby direct UV photolysis in the H2O2/UV treatment. Otherthan that, degradation mechanisms of these contrastmedia by ozonation and AOPs are still largely unknown;kinetic data are very scarce and no degradation by-product was identified. It is probably desirable to discernthe identity and property of degradation by-products oftriiodinated X-ray contrast media because even thoughthe parent compounds are non-toxic (Steger-Hartmannet al., 1999), their degradation products may have somepotential ecological and public health impacts. Finally,there are two ionic triiodinated X-ray contrast media,iothalamic acid and ioxithalamic acid, that have notbeen studied for their degradation by ozonation or AOPbut been detected in the aquatic environment (Ternes andHirsch, 2000).

HO

OHHN

OI

I

I

NO

OH

NHO

OH

HO

iomeprol (777.09)

HNO

I

I

I

NH

NHO

OH

HO

OH

HOO OH

iopamidol (777.09)

HNO

I

I

I

NH

NO O

iopromide (791.12)

O

HO

HO

OHHO

HNO

I

I

I

N

NHO O

HO

HO

OHHO

OOH

iopentol (835.17)

HOO

I

I

I

NH

NH OO

diatrizoate (613.91)

FIGURE 36. X-ray contrast media.

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CONCLUDING REMARKS

Pharmaceutical contamination of surface water andgroundwater is an emerging issue in environmentalscience and engineering. After the administration tohumans and animals, these medical drugs and non-ther-apeutic agents are partially metabolized and excreted inthe urine and/or the feces, and subsequently enter theaquatic environment through a number of routes (seeFigure 1). Some of the pharmaceuticals are fairly biode-gradable, while others are more persistent and mobile inthe aquatic environment. A large number of pharmaceu-ticals have been detected in domestic sewage treatmentplant effluents, hospital wastewaters, surface water, andgroundwater at ng/L to mg/L levels. Although there is noclear evidence of immediate public health impacts of thesetrace pharmaceuticals in water, there are several groupsof substances with unambiguous toxic and estrogenicproperties such as antibiotics, cytostatic agents, and nat-ural and synthetic hormones, which can indeed affectpopulations of aquatic organisms. Besides, the chronichealth effects of a mixture of biologically active sub-stances are still largely unknown. Therefore, the removalof these substances before entering the aquatic environ-ment and drinking water is probably desirable based onthe precautionary principle. Chemical oxidation by ozoneand AOPs has been thought to be one of the promisingtreatment methods to deal with this problem.

The studies reviewed here have indeed showed satisfac-tory efficiencies of ozonation and AOPs for the degradationof a wide range of pharmaceutical compounds in aqueoussolution. Some pharmaceuticals are extremely reactivetoward molecular ozone, including some antibiotics, ananticonvulsant carbamazepine, a NSAID diclofenac, andan estrogen 17b-estradiol. These pharmaceuticals can becharacterized by one or more functional groups and moi-eties in the molecules such as non-aromatic carbon-carbondouble bonds, amines, thioether sulfurs, and activated aro-matic rings. There are also some pharmaceuticals relativelyresistant to ozonation, including a lipid regulator metabo-lite clofibric acid, an anti-anxiety agent diazepam, and aNSAID ibuprofen. Triiodinated X-ray contrast media,which are biologically stable, non-therapeutic medicalagents, are particularly refractory to ozonation. Ozone-based AOPs, Fenton-type processes and photochemicalAOPs are generally more effective than ozonation aloneprimarily due to enhanced generation of hydroxyl radicalsand photon-initiated cleavage of carbon-halogen bonds,thus are recommended for the treatment of these recalci-trant substances. There are a number of pharmaceuticalsthat have not been studied for their degradation by ozona-tion or AOP but been found in the aquatic environment(Table 3). These pharmaceuticals may warrant furtherstudy.

The degree of degradation of water and wastewaterpollutants, including pharmaceutical compounds, achieved

by ozonation and AOPs depends on a number of factors.Oxidant dose, concentration of pharmaceuticals, variouswater quality parameters, and mode of operation are a fewexamples. It has been demonstrated that the presence ofinorganic and organic water and wastewater constituentsoften inhibit the degradation of target pollutants throughoxidant competition and/or radical scavenging mechan-isms, which has resulted in higher oxidant requirements.In addition to the oxidative transformation of parent phar-maceuticals, certain levels of mineralization have beendemonstrated by ozonation and AOPs. It should benoted that complete mineralization may not be an absoluterequirement in both water and wastewater treatment. Inwastewater treatment, improvement of biodegradability isperhaps the more important goal to be achieved, especiallyfor the pharmaceuticals that are resistant to biodegrada-tion, such as antibiotics, cytostatic agents, hormones,X-ray contrast agents, carbamazepine, and some acidicdrugs like clofibric acid. However, the achievement ofthis goal has been confirmed only for some antibioticsthrough limited types of treatment process (seeAppendix). Likewise, toxicity reduction by ozonation orAOPs has been evaluated in just a handful of cases, exclud-ing the estrogenicity of natural and synthetic hormones. Inwater treatment, considering the very small amounts ofpharmaceuticals in finished drinking water, the destructionof key molecular structures for biological activity may besufficient to protect public health, if the degradation by-products do not exert similar or other unprecedented andundesirable effects. Thus, identification and characteriza-tion of degradation by-products and intermediates ishighly important task, although this has not been doneyet for all the substances discussed here. In addition,kinetic data are still scarce for many pharmaceuticalsreviewed here. Further validation may be as well requiredfor those substances already studied for their degradationkinetics by ozone or hydroxyl radicals, because of somediscrepancies, high experimental errors, or limited pH andtemperature coverage.

In order to minimize the load of pharmaceuticals in theaquatic environment, a number of actions should betaken simultaneously, besides the end-of-pipe treatmentapproach discussed in this review. It has been recognizedthat source separation is one of the key for a successfulwaste management practice in general (Jonsson et al.,1997; Tanskanen et al., 1998), and this approach shouldbe applicable to wastewater pollutants such as pharma-ceuticals. As many pharmaceuticals are excreted mostlyin the urine, its separation from the solids portion (feces)of wastewater and the provision for separate treatmentsystem may be of an ultimate solution particularly forhighly contaminated hospital wastewater (Larsen et al.,2004). Ozonation and advanced oxidation are likely sui-table for the treatment of the pharmaceutical-laden urine,as suggested by Larsen et al. (2004); thus this strategyshould be investigated in the future.

388 K. Ikehata et al. December 2006

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ACKNOWLEDGMENT

The authors greatly appreciate the financial supportprovided by Alberta Ingenuity Fund (AIF) and AlbertaIngenuity Centre for Water Research (AICWR).

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Pharmaceutical Class Occurrences

Ciprofloxacin Antibiotics, fluoroquinolone Up to 0.4 mg/L in final effluents from WWTP(Miao et al., 2004)

Norfloxacin Antibiotics, fluoroquinolone Up to 0.112 mg/L in final effluents from WWTP(Miao et al., 2004), up to 0.12 mg/L in surface water(Kolpin et al., 2002)

Clindamycin Antibiotics, lincosamide Up to 1.1 mg/L in surface water (Batt and Aga, 2005),up to 32 ng/L in surface water (Christian et al., 2003)

Doxycycline Antibiotics, tetracycline Up to 0.046 mg/L in WWTP (Miao et al., 2004)Oxytetracycline Antibiotics, tetracycline Up to 0.34 mg/L in surface water (Kolpin et al., 2002)Paroxetine Antidepressant Its environmental fate was studied (Kwon and Armbrust,

2004), a low risk to environment (Cunningham et al., 2004)Ketoprofen Non-steroidal anti-inflammatory

drugUp to 0.38 and 0.12 mg/L in effluents from WWTP andsurface water, respectively (Ternes, 1998), up to 0.2mg/L in effluents from WWTP (Soulet et al., 2002)

Mefenamic acid Non-steroidal anti-inflammatorydrug

Up to 0.6 mg/L in effluents from WWTP (Soulet et al., 2002)

Belomycin Cytostatic drug Up to 19, 17, and 13 ng/L in effluents from WWTP,surface water, and potable water, respectively(Halling-Sørensen et al., 1998)

Equilin Steroid hormone, estrogen Up to 0.147 mg/L in surface water (Kolpin et al., 2002)Testosterone Steroid hormone, androgen Up to 0.214 mg/L in surface water (Kolpin et al., 2002)Iothalamic acid Triiodinated X-ray contrast media Up to 0.049 mg/L in groundwater (Ternes and Hirsch, 2000)Ioxithalamic acid Triiodinated X-ray contrast media Up to 0.010 mg/L in groundwater (Ternes and Hirsch, 2000)

Note: WWTP=wastewater treatment plant.

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PE

ND

IX.

Sum

mary

of

the

Degra

datio

nof

Pharm

aceutic

als

by

Ozonatio

nand

AO

Ps

Pharm

aceutical

(CAS#)

Class

Applied

process

Degreeofdegradation

Kinetic

parameters

Interm

ediates/

by-products

Biodegradabilityand

toxicity

Amoxicillin

(61336-70-7)

Antibiotic,

b-lactam

Ozonation

Complete

conversionof

400mg/L

amoxicillin

(Arslan-A

latonand

Dogruel,2004),86%

ofinitialCOD

(=1,395mg/L)

removed

atpH

11.5

(buffered)from

filtered

form

ulation

wastew

ater.49%

of

initialCOD

and52%

ofinitialTOC

(=920

mg/L)removed

innon-buffered

system

(Arslan-A

latonand

Dogruel,2004),81%

COD

removalby

simulatedchem

ical-

biologicaltreatm

ent

(Arslan-A

latonet

al.,

2004)

kO3=

4·103M�1�s�

1

atpH

2.5,6

·106

M�1�s�

1atpH

7(A

ndreozzi

etal.,2005)

Elementalsulfur

(Arslan-A

laton

andDogruel,

2004),

hydroxylationof

phenolring

(Andreozziet

al.,

2005)

BOD

5increasedfrom

zero

to60mg/L

atpH

11.5

(non-buffered

)(A

rslan-A

latonand

Dogruel,2004),

BOD

5/C

OD

increased

from

zero

to0.08

(Arslan-A

latonand

Dogruel,2004),to

0.37

(Arslan-A

latonet

al.,

2004)

O3/H

2O

283%

ofinitialCOD

(=830mg/L)removed

atpH

10.5,72%

COD

removalbysimulated

chem

ical-biological

treatm

ent(A

rslan-

Alatonet

al.,2004)

N/D

N/D

BOD

5/C

OD

increased

from

zero

to0.45

(Arslan-A

latonet

al.,

2004)

H2O

2/U

V22%

ofinitialCOD

(=1,395mg/L)and

6%

ofinitialTOC

(=920mg/L)removed

atpH

7([H

2O

2] 0=

1.02g/L)

(Arslan-A

latonand

Dogruel,2004)

k�O

H=

3.93

·109

M�1�s�

1atpH

5.5

(Andreozzi

etal.,2005)

N/D

BOD

5increasedfrom

zero

to10mg/L

(Arslan-A

latonand

Dogruel,2004)

(Contin

ued)

Ozonation and Advanced Oxidation of Pharmaceuticals December 2006 393

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2007

Fenton/

Fenton-like

Complete

conversionof

400mg/L

amoxicillin

(Arslan-A

latonand

Dogruel,2004),61%

ofinitialCOD

(=1,395mg/L)and

33%

ofinitialTOC

(=920mg/L)removed

byFenton,46%

COD

and18%

TOC

by

Fenton-like,

both

at

pH=

3,1mM

Fe2

+

orFe3

+,20mM

H2O

2

(Arslan-A

latonand

Dogruel,2004)

N/D

N/D

BOD

5increasedfrom

zero

to6mg/L

(Arslan-A

latonand

Dogruel,2004)

Photo-Fenton/

Photo-

Fenton-like

56%

ofinitialCOD

(=1,395mg/L)and

46%

ofinitialTOC

(=920mg/L)removed

byphoto-Fenton,66%

COD

and42%

TOC

removed

byphoto-

Fenton-like,

both

at

pH

3,1mM

Fe2

+or

Fe3

+,20mM

H2O

2

(Arslan-A

latonand

Dogruel,2004)

N/D

N/D

BOD

5increasedfrom

zero

to4mg/L

(photo-

Fenton)and21mg/L

(photo-F

enton-like)

(Arslan-A

latonand

Dogruel,2004)

Photolysis

NoCOD/TOC

reduction(A

rslan-

AlatonandDogruel,

2004)

FP=

0.571

mol�E

instein�1atpH

5.5

(Andreozziet

al.,

2005)

N/D

NoBOD

5increase

(Arslan-A

latonand

Dogruel,2004)

PenicillinVK

(132-98-9)

Antibiotic,

b-lactam

Ozonation

70%

ofinitialCOD

(=450mg/L)and

40%

ofinitialTOC

(=162mg/L)removed

atpH

7-11

(Akmehmet

Balcio

� glu

andOtker,2003)

N/D

N/D

BOD

5/C

OD

increased

from

0to

0.25

(Akmehmet

Balcio

� glu

andOtker,2003)

AP

PE

ND

IX.

(Contin

ued)

Pharm

aceutical

(CAS#)

Class

Applied

process

Degreeofdegradation

Kinetic

parameters

Interm

ediates/

by-products

Biodegradabilityand

toxicity

394 K. Ikehata et al. December 2006

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2007

O3/H

2O

295%

ofinitialCOD

(=450mg/L)removed

atpH

7(O

3:H

2O

2

molarratio=

3.08:1)

(Akmehmet

Balcio

� glu

andOtker,2003)

N/D

N/D

N/D

PenicillinG

(61-33-6)

Antibiotic,

b-lactam

Ozonation

50%

ofinitialCOD

(=600mg/L)and

50%

ofTOC

(=226

mg/L)removed

atpH

12(A

rslan-A

latonand

Caglayan,2005).

kCOD,O

3=

0.67M�1�s�

1

atpH

7,0.042M�1�s�

1

atpH

3(A

rslan-

AlatonandCaglayan,

2005)

N/D

N/D

Photo-Fenton-

like

56%

ofinitialCOD

(=600mg/L)and

42%

ofinitialTOC

(=226mg/L)removed

atpH

3(A

rslan-

AlatonandGurses,

2004)

N/D

N/D

Daphnia

magnaacute

toxicityreduced,

BOD

5/C

OD

increased

from

0.25to

0.45

(Arslan-A

latonand

Gurses,2004)

Sultamicillin

(76497-13-7)

Antibiotic,

b-lactam

Ozonation-

biodegradation

33%

ofinitialCOD

(=710mg/L)and

24%

ofinitialTOC

(=200mg/L)removed

atpH

11(C

okgor

etal.,2004).

N/D

N/D

BOD

5/C

OD

increasedfrom

0.02to

0.27

(Cokgoret

al.,

2004)

Cefradine

(38821-53-3)

Antibiotic,

b-lactam

(cephalosporin)

TiO

2/hv

Complete

conversionof

70mg/L

cefradine,

additionofH

2O

2

accelerate

decomposition(Fan

etal.,2002)

N/D

N/D

N/D

(Contin

ued)

Ozonation and Advanced Oxidation of Pharmaceuticals December 2006 395

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2007

Ceftriaxone

(73384-59-5)

Antibiotic,

b-lactam

(cephalosporin)

Ozonation

53%,74%

and82%

of

initialCOD

(=450

mg/L)removed

atpH

3,7,and11

(Akmehmet

Balcio

� glu

andOtker,2003)

N/D

N/D

BOD

5/C

OD

increased

from

zero

to0.10

(Akmehmet

Balcio

� glu

andOtker,2003)

O3/H

2O

2About90%

ofinitial

COD

(=450mg/L)

removed

atpH

7(A

kmehmet

Balcio

� glu

andOtker,2003)

N/D

N/D

N/D

MMTD

(29490-19-5)

Interm

ediate

of

cefazolin

(cephalosporin

antibiotic)

H2O

2/U

VComplete

conversionof

1mg/L

MMTD

in20

min,60%

TOC

removed

in4h(Lopez

etal.,2002)

k�O

H=

1.6

·1010

M�1�s�

1at25� C

(Lopez

etal.,2003)

Degradation

pathway

proposed(Lopez

etal.,2002)

N/D

Photolysis

About90%

conversion

of1mg/L

MMTD

in60min,no

mineralization(Lopez

etal.,2002)

FP=

12

mmol�E

instein�1

(Lopez

etal.,2002)

Oneby-product

identified

(Lopez

etal.,2002)

N/D

MMTD-M

e(1925-78-6)

Interm

ediate

of

cefazolin

(cephalosporin

antibiotic)

H2O

2/U

VComplete

conversion

of1mg/L

MMTD-M

ein

20min,80%

TOC

removed

in4h(Bozzi

etal.,2002)

k�O

H=

8.3

·108

M�1�s�

1

at25� C

(Lopez

etal.,2003)

S-O

xidation

precedes

mineralization

(Bozziet

al.,

2002)

N/D

photolysis

Complete

conversion

of1mg/L

MMTD-M

ein

60min,no

mineralization(Bozzi

etal.,2002)

FP=

14.1

mmol�E

instein�1

(Lopez

etal.,2003)

Twoby-products

identified

(Bozzi

etal.,2002)

N/D

Azithromycin

(83905-01-5)

Antibiotic,

macrolide

Ozonation

Complete

conversionof

azithromycin

(concentration

unknown)atpH

7(H

uber

etal.,2005)

N/D

N/D

N/D

AP

PE

ND

IX.

(Contin

ued)

Pharm

aceutical

(CAS#)

Class

Applied

process

Degreeofdegradation

Kinetic

parameters

Interm

ediates/

by-products

Biodegradabilityand

toxicity

396 K. Ikehata et al. December 2006

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2007

Clarithromycin

(81103-11-9)

Antibiotic,

macrolide

Ozonation

Complete

conversionof

0.21-2

mg/L

clarithromycinatpH

7-7.2

(Ternes

etal.,

2003;Huber

etal.,

2005)

N/D

N/D

N/D

Erythromycin

(114-07-8)

Antibiotic,

macrolide

Ozonation

Complete

conversionof

0.62-2

mg/L

(dehydro)erythromycin

atpH

7-7.2

(Ternes

etal.,2003;

Huber

etal.,2005)

N/D

N/D

N/D

Lincomycin

(154-21-2)

Antibiotic,

lincosamide

Ozonation

N/D

kO3=

3.26

·105

M�1�s�

1(protonated),

2.43

·106M�1�s�

1

(neutral)(Q

ianget

al.,

2004)

N/D

N/D

TiO

2/hn

Complete

conversionof

50mg/L

lincomycinin

2hatpH

6,60%

TOC

removalin

5h

(Addamoet

al.,2005)

Langmuir–Hinshelwood

kinetic

model

developed

(Addamo

etal.,2005)

Sulfate

ion

(Addamoet

al.,

2005)

N/D

Roxithromycin

(80214-83-1)

Antibiotic,

macrolide

Ozonation

Complete

conversionof

0.54-2

mg/L

roxithromycinatpH

7-7.2

(Ternes

etal.,

2003;Huber

etal.,

2005)

kO3=

4.5

·106M�1�s�

1

atpH

>8.8

(Huber

etal.,2003)

N/D

N/D

Enrofloxacin

(93106-60-6)

Antibiotic,

quinolone

(veterinary)

Ozonation

88%

ofinitialCOD

(=450mg/L)and

50%

ofinitialTOC

(=165mg/L)removed

atpH

7(A

kmehmet

Balcio

� gluandOtker,

2003),successfully

decontaminated

enrofloxacin-loaded

zeolite

(Otker

and

Akmehmet-B

alcio

� glu,

2005)

N/D

N/D

BOD

5/C

OD

increased

from

0.07to

0.38

(Akmehmet

Balcio

� glu

andOtker,2003)

(Contin

ued)

Ozonation and Advanced Oxidation of Pharmaceuticals December 2006 397

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2007

Ofloxacin

(levofloxacin;

637328-10-0)

Antibiotic,

quinolone

Ozonation

Complete

conversionof

560mg

/Lofloxacinin

amixture

ofsix

pharm

aceuticalsatpH

7.4

(Andreozziet

al.,

2004)

N/D

N/D

Algalandprotozoan

toxicityelim

inated(as

amixture)(A

ndreozzi

etal.,2004)

H2O

2/U

VComplete

conversionof

560mg

/Lofloxacinin

amixture

ofsix

pharm

aceuticalsatpH

7.4

(Andreozziet

al.,

2004)

N/D

N/D

Algaltoxicity

elim

inated,protozoan

toxicityreduced(asa

mixture)(A

ndreozzi

etal.,2004)

TiO

2/hn

Incomplete

conversion

of560mg

/Lofloxacin

inamixture

ofsix

pharm

aceuticals

(Andreozziet

al.,

2004)

N/D

N/D

Notoxicityreduction

(asamixture)

(Andreozziet

al.,

2004)

Sulfachlorpyridazine

(80-32-0)

Antibiotic,

sulfonamide

Ozonation

Complete

conversionof

50mg

/Lsulfachlorpyridazineat

pH

7.5

in1.5

min

(Adamset

al.,2002)

N/D

N/D

N/D

Fenton

N/D

(See

Table

2)

N/D

N/D

Sulfadiazine

(68-35-9)

Antibiotic,

sulfonamide

Ozonation

Complete

conversionof

2mg

/Lsulfadiazineat

pH

7(H

uber

etal.,

2005)

N/D

N/D

N/D

Fenton

N/D

(See

Table

2)

N/D

N/D

TiO

2/hn

80%

conversionof15

mg/L

sulfadiazinein

30min,80%

and15%

recoveryofsulfurand

nitrogen,respectively

(Calzaet

al.,2004b)

N/D

SO

2�

4;N

Hþ 4;

NO� 3;hydroxy-

latedsulfadiazine

(Calzaet

al.,

2004b)

N/D

AP

PE

ND

IX.

(Contin

ued)

Pharm

aceutical

(CAS#)

Class

Applied

process

Degreeofdegradation

Kinetic

parameters

Interm

ediates/

by-products

Biodegradabilityand

toxicity

398 K. Ikehata et al. December 2006

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2007

Sulfadim

ethoxine

(122-11-2)

Antibiotic,

sulfonamide

Ozonation

Complete

conversionof

50mg

/Lsulfadim

ethoxineat

pH

7.5

in1.5

min

(Adamset

al.,2002)

N/D

N/D

N/D

Fenton

N/D

(See

Table

2)

N/D

N/D

TiO

2/hn

Complete

conversionof

15mg/L

sulfadim

ethoxinein

30

min,90%

and70%

recoveryofsulfurand

nitrogen,respectively,

in1h(C

alzaet

al.,

2004b)

N/D

SO

2�

4;N

Hþ 4;

NO� 3;tw

otypes

ofhydroxylated

sulfadim

ethoxine,

2,6-dim

ethoxy-4-

aminopyrimidine

(Calzaet

al.,

2004b)

N/D

Sulfamerazine

(127-79-7)

Antibiotic,

sulfonamide

Ozonation

Complete

conversionof

50mg

/Lsulfamerazine

atpH

7.5

in1.5

min

(Adamset

al.,2002)

N/D

N/D

N/D

Fenton

N/D

(See

Table

2)

N/D

N/D

TiO

2/hn

85%

conversionof15

mg/L

sulfamerazinein

30min,70%

and18%

recoveryofsulfurand

nitrogen,respectively

in30min

(Calzaet

al.,

2004b)

N/D

SO

2�

4;N

Hþ 4;

NO� 3;hydroxy-

latedsulfamera-

zine,

4-m

ethyl-2-

aminopyrimidine

(Calzaet

al.,

2004b)

N/D

Sulfamethazine

(57-68-1)

Antibiotic,

sulfonamide

Ozonation

Complete

conversionof

50mg

/Lsulfamethazo

nie

atpH

7.5

in1.5

min

(Adams

etal.,2002),

N4-acetylationlower

thereactivitytoward

ozonation(H

uber

etal.,2005)

N/D

N/D

N/D

Fenton

N/D

(See

Table

2)

N/D

N/D

(Contin

ued)

Ozonation and Advanced Oxidation of Pharmaceuticals December 2006 399

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2007

Sulfamethoxazole

(723-46-6)

Antibiotic,

sulfonamide

Ozonation

Complete

conversionof

0.62-2

mg/L

sulfamethoxazole

at

pH

7-7.2

(Ternes

etal.,

2003;Huber

etal.,

2005),Complete

conversionof2.24mg/

Lsulfamethoxazole

inamixture

ofsix

pharm

aceuticalsatpH

7.4

(Andreozziet

al.,

2004),N

4-acetylation

reducesreactivity

(Huber

etal.,2005)

kO3=

2.5

·106M�1�s�

1

at20� C

(Huber

etal.,

2003)

N/D

Algalandprotozoan

toxicityelim

inated(as

amixture)(A

ndreozzi

etal.,2004)

H2O

2/U

VComplete

conversionof

2.24mg/L

sulfamethoxazole

ina

mixture

ofsix

pharm

aceuticalsatpH

7.4

(Andreozziet

al.,

2004)

k�O

H=

5.5

·109

M�1�s�

1atpH

7and

25� C

(Huber

etal.,

2003),seealsoTable

2

N/D

Algaltoxicity

elim

inated,protozoan

toxicityreduced(asa

mixture)(A

ndreozzi

etal.,2004)

TiO

2/hn

Incomplete

conversion

of2.24mg/L

sulfamethoxazole

ina

mixture

ofsix

pharm

aceuticals

(Andreozziet

al.,

2004)

N/D

N/D

Notoxicityreduction

(asamixture)

(Andreozziet

al.,

2004)

Sulfamethizole

(144-82-1)

Antibiotic,

sulfonamide

Fenton

N/D

(See

Table

2)

N/D

N/D

Sulfamoxole

(729-99-7)

Antibiotic,

sulfonamide

Fenton

Unstable

inweakly

acidicsolution(Boreen

etal.,2004)

(See

Table

2)

N/D

N/D

Sulfapyridine

(144-83-2)

Antibiotic,

sulfonamide

Ozonation

Complete

conversionof

2mg

/Lsulfapyridineat

pH

7(H

uber

etal.,

2005)

N/D

N/D

N/D

AP

PE

ND

IX.

(Contin

ued)

Pharm

aceutical

(CAS#)

Class

Applied

process

Degreeofdegradation

Kinetic

parameters

Interm

ediates/

by-products

Biodegradabilityand

toxicity

400 K. Ikehata et al. December 2006

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2007

Sulfathiazole

(72-14-0)

Antibiotic,

sulfonamide

Ozonation

Complete

conversionof

50mg

/Lsulfadim

ethoxineat

pH

7.5

in1.5

min

(Adamset

al.,2002),

Complete

conversion

of2mg

/Lsulfathiazole

atpH

7(H

uber

etal.,

2005)

N/D

N/D

N/D

Fenton

N/D

(See

Table

2)

N/D

N/D

TiO

2/hn

Complete

conversionof

15mg/L

sulfathiazole

in30min,quantitative

recoveryofsulfur,

90%

recoveryof

nitrogen

(Calzaet

al.,

2004b)

N/D

SO

2�

4;N

Hþ 4;

NO� 3,hydroxy-

latedsulfathia-

zole

(Calzaet

al.,

2004b)

N/D

Sulfisoxazole

(127-69-5)

Antibiotic,

sulfonamide

Fenton

N/D

(See

Table

2)

N/D

N/D

Carbadox

(6804-07-5)

Antibiotic,

veterinary

Ozonation

Complete

conversionof

50mg

/Lcarbadoxat

pH

7.5

in1.5

min

(Adamset

al.,2002)

N/D

N/D

N/D

Spectinomycin

(1695-77-8)

Antibiotic,

miscellaneous

Ozonation

N/D

kO3=

1.27

·106

M�1�s�

1(neutral),3.30

·105M�1�s�

1(m

ono-

protonated)(A

dams

etal.,2002)

N/D

N/D

Tetracycline

(60-54-8)

Antibiotic,

tetracycline

TiO

2/hn(+

direct

photolysis)

Complete

conversionof

50mg/L

tetracyclinein

2h,90%

TOC

removalin

6h(D

iPaola

etal.,2004;

Addamoet

al.,2005)

Kinetic

model

presented

(Addamoet

al.,2005)

N/D

N/D

Trimethoprim

(738-70-5)

Antibiotic,

chem

otherapeu

tic

Complete

conversionof

0.34-50mg

/Ltrim

ethoprim

at

pH

7-7.5

(Adams

etal.,2002;

Ternes

etal.,

2003)

N/D

(Contin

ued)

Ozonation and Advanced Oxidation of Pharmaceuticals December 2006 401

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2007

N/D

N/D

Buspirone

(36505-84-7)

Anti-anxiety

agent

TiO

2/hn

Complete

conversionof

15mg/L

buspironein

30min

(Calzaet

al.,

2004a)

N/D

Degradation

pathway

proposed(C

alza

etal.,2004a)

N/D

Carbamazepine

(298-46-4)

Anticonvulsant

Ozonation

Complete

conversionof

0.5-118mg/L

carbamazepine

(Andreozziet

al.,

2002;Ternes

etal.,

2002;Huber

etal.,

2003),30%

mineralizationasCO

2

(Andreozziet

al.,

2002)

kO3=

7.81

·104

M�1�s�

1(A

ndreozzi

etal.,2002),3

·105

M�1�s�

1(H

uber

etal.,

2003)

Degradation

pathway

proposed

(Andreozziet

al.,

2002;McD

owell

etal.,2005)

Algaltoxicity

dim

inished

(Andreozzi

etal.,2002)

H2O

2/U

VComplete

conversionof

4.7

mg/L

carbamazepineand

35%

TOC

removalin

4min

atpH

5.0

(Vognaet

al.,2004a)

k�O

H=

2.05

·109

M�1�s�

1(V

ognaet

al.,

2004a),8.8

·109

M�1�s�

1(H

uber

etal.,

2003)

Degradation

pathway

proposed(V

ogna

etal.,2004a)

N/D

TiO

2/hn

>90%

conversionof4.2

mg/L

carbamazepine

in9min

(Dolland

Frimmel,2005c),pilot

scale

treatm

ent(D

oll

andFrimmel,2005a)

Pseudo-1st-order

kinetics(D

olland

Frimmel,2004)

Degradation

pathway

proposed(D

oll

andFrimmel,

2005b)

N/D

Diazepam

(439-14-5)

Anti-anxiety

agent/

anticonvulsant

Ozonation

24%-65%

conversionof

142mg

/Ldiazepam

at

pH

9and10� C

(Huber

etal.,2003)

kO3=

0.75M�1�s�

1at

20� C

(Huber

etal.,

2003)

N/D

N/D

H2O

2/U

V,g-

radiolysis

N/D

k�O

H=

7.2

·109

M�1�s�

1atpH

7and

25� C

(Huber

etal.,

2003)

N/D

N/D

Primidone

(125-33-7)

Anticonvulsant

Ozonation

Upto

87%

conversion

of1mg

/Lprimidoneat

pH

7.8

and23� C

(Ternes

etal.,2002)

N/D

N/D

N/D

AP

PE

ND

IX.

(Contin

ued)

Pharm

aceutical

(CAS#)

Class

Applied

process

Degreeofdegradation

Kinetic

parameters

Interm

ediates/

by-products

Biodegradabilityand

toxicity

402 K. Ikehata et al. December 2006

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2007

Diclofenac

(15307-86-5)

Non-steroidal

anti-

inflammatory

drug(N

SAID

)

Ozonation

Complete

conversionof

1mg

/L–296mg/L

diclofenac(Zwiener

andFrimmel,2000;

Ternes

etal.,2002;

Ternes

etal.,2003;

Vognaet

al.,2004b;

Huber

etal.,2005),

>30%

TOC

removal,

95%

dechlorination

(Vognaet

al.,2004b)

kO3=

1·106M�1�s�

1

at20� C

(Huber

etal.,

2003),1.84

·104

M�1�s�

1at25� C

(Vognaet

al.,2004b)

Degradation

pathway

proposed(V

ogna

etal.,2004b)

Algalandprotozoan

toxicityelim

inated(as

amixture)(A

ndreozzi

etal.,2004)

O3/H

2O

2Complete

conversionof

2mg

/Ldiclofenacin

10

min

(Zwiener

and

Frimmel,2000)

N/D

N/D

N/D

H2O

2/U

VComplete

conversionof

296mg/L

diclofenac,

40%

TOC

removal,

50%

dechlorination

(Vognaet

al.,2004b)

N/D

Degradation

pathway

proposed(V

ogna

etal.,2004b)

Algaltoxicity

elim

inated,protozoan

toxicityreduced(asa

mixture)(A

ndreozzi

etal.,2004)

g-radiolysis

N/D

k�O

H=

7.2

·109

M�1�s�

1atpH

7and

25� C

(Huber

etal.,

2003)

N/D

N/D

Fenton

NoconversionatpH

3.5

and22� C

(Packer

etal.,2003)

N/D

N/D

N/D

Photo-Fenton

Complete

conversionof

10-80mg/L

diclofenac,

>90%

TOC

removal

atpH

2.8

and50� C

(Ravinaet

al.,2002),

poorperform

ance

inbuffered

acidic

and

neutralmedia

(Perez-

Estradaet

al.,2005a)

N/D

Degradation

pathway

proposed(Perez-

Estradaet

al.,

2005b)

N/D

TiO

2/hn

Complete

conversionof

50mg/L

diclofenac,

55%

DOC

removal

(Perez-Estradaet

al.,

2005a)

N/D

N/D

Notoxicityreduction

(asamixture)

(Andreozziet

al.,

2004)

(Contin

ued)

Ozonation and Advanced Oxidation of Pharmaceuticals December 2006 403

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2007

Ibuprofen

(15687-27-1)

Non-steroidal

anti-

inflammatory

drug(N

SAID

)

Ozonation

Lessreactivethanother

NSAID

s(Zwiener

and

Frimmel,2000;Ternes

etal.,2003;Huber

etal.,2005)

kO3=

9.6

M�1�s�

1at20

� C(H

uber

etal.,2003)

N/D

N/D

O3/H

2O

2Complete

conversionof

2mg

/Ldiclofenacat

pH

7.5

and10� C

(Zwiener

and

Frimmel,2000)

N/D

N/D

N/D

H2O

2/U

VN/D

k�O

H=

7.4

·109

M�1�s�

1atpH

7and

25� C

(Huber

etal.,

2003)

N/D

N/D

Fenton

N/D

k�O

H=

6.5

·109

M�1�s�

1atpH

3.5

and

22� C

(Packer

etal.,

2003)

N/D

N/D

Indomethacin

(53-86-1)

Non-steroidal

anti-

inflammatory

drug(N

SAID

)

Ozonation

Complete

conversionof

0.1

mg/L

indomethacin

atpH

7.2

(Ternes

etal.,2003)

N/D

N/D

N/D

Naproxen

(22204-53-1)

Non-steroidal

anti-

inflammatory

drug(N

SAID

)

Ozonation

Complete

conversionof

0.1

mg/L

naproxen

at

pH

7.2

(Ternes

etal.,

2003)

N/D

N/D

N/D

Fenton

N/D

k�O

H=

9.6

·109

M�1�s�

1atpH

3.5

and

22� C

(Packer

etal.,

2003)

N/D

N/D

AP

PE

ND

IX.

(Contin

ued)

Pharm

aceutical

(CAS#)

Class

Applied

process

Degreeofdegradation

Kinetic

parameters

Interm

ediates/

by-products

Biodegradabilityand

toxicity

404 K. Ikehata et al. December 2006

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2007

Salicylicacid

(69-72-7)

Adecomposition

product

of

acetylsalicylic

acid

Ozonation

60%

conversionof0.65

mg/L

salicylicacid

(Khanet

al.,2004)

N/D

N/D

N/D

Paracetamol

(103-90-2)

Antipyretic/

analgesic

Ozonation

Complete

conversionof

0.8

g/L

paracetamolat

pH

2and7in

20min,

20%

and30%

TOC

removalatpH

2and

7,respectively,in

2h

(Andreozziet

al.,

2003a)

kO3=

1.41

·103

M�1�s�

1(neutral),9.91

·108M�1�s�

1

(dissociated)

(Andreozziet

al.,

2003a)

Degradation

pathway

proposed

(Andreozziet

al.,

2003a)

N/D

H2O

2/U

V>90%

conversionof

1.51mg/L

paracetamolatpH

5.5

in1min,40%

TOC

removalin

4min

(Andreozziet

al.,

2003a)

k�O

H=

2.2

·109

M�1�s�

1(A

ndreozzi

etal.,2003a)

Degradation

pathway

proposed(V

ogna

etal.,2002;

Andreozziet

al.,

2003a)

N/D

Anodic

oxidation

usinga

boron-doped

diamond

anode

Complete

conversion

and70%-98%

TOC

removalof78-948mg/

Lparacetamol(Brillas

etal.,2005)

N/D

Oxalicacid,oxamic

acid(Brillas

etal.,2005)

N/D

Atenolol

(29122-68-7)

b-blocker

Ozonation

Complete

conversionof

0.36mg

/Latenololat

pH

7.2

(Ternes

etal.,

2003)

N/D

N/D

N/D

Celiprolol

(56980-93-9)

b-blocker

Ozonation

Complete

conversionof

0.28mg

/Lceliprololat

pH

7.2

(Ternes

etal.,

2003)

N/D

N/D

N/D

Metoprolol

(37350-58-6)

b-blocker

Ozonation

Complete

conversionof

1.7

mg/L

metoprololat

pH

7.2

(Ternes

etal.,

2003)

N/D

N/D

N/D

Propranolol

(525-66-6)

b-blocker

Ozonation

Complete

conversionof

0.18-325.5

mg/L

propranololatpH

7.2-

7.4

(Ternes

etal.,2003;

Andreozziet

al.,2004)

N/D

N/D

Algalandprotozoan

toxicityelim

inated(as

amixture)(A

ndreozzi

etal.,2004)

(Contin

ued)

Ozonation and Advanced Oxidation of Pharmaceuticals December 2006 405

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2007

H2O

2/U

VComplete

conversionof

325.5

mg/L

propranololatpH

7.4

(Andreozziet

al.,

2004)

N/D

N/D

Algaltoxicity

elim

inated,protozoan

toxicityreduced(asa

mixture)(A

ndreozzi

etal.,2004)

TiO

2/hn

Incomplete

conversion

of325.5

mg/L

propranolol

(Andreozziet

al.,

2004)

N/D

N/D

Notoxicityreduction

(asamixture)

(Andreozziet

al.,

2004)

Sotalol(3930-20-9)

b-blocker

Ozonation

Complete

conversionof

1.32mg

/Lsotalolat

pH

7.2

(Ternes

etal.,

2003)

N/D

N/D

N/D

Aclarubicin

(57576-44-0)

Cytostaticdrug,

anthracycline

Fenton

Complete

conversionof

0.5

g/L

aclarubicin

(Castegnaro

etal.,

1997),H

2O

2alonewas

less

effective

(Castegnaro

etal.,

1997)

N/D

Absence

of

fluorescentby-

products

(Castegnaro

etal.,1997)

Noresidual

mutagenicity

(Castegnaro

etal.,

1997)

Daunorubicin

(20830-81-3)

Cytostaticdrug,

anthracycline

Fenton

Complete

conversionof

5g/L

daunorubicin

(Castegnaro

etal.,

1997),H

2O

2alonewas

less

effective

(Castegnaro

etal.,

1997)

N/D

Absence

of

fluorescentby-

products

(Castegnaro

etal.,1997)

Noresidual

mutagenicity

(Castegnaro

etal.,

1997)

Doxorubicin

(23214-92-8)

Cytostaticdrug,

anthracycline

Fenton

Complete

conversionof

0.4

g/L

doxorubicin

(Castegnaro

etal.,

1997),H

2O

2alonewas

less

effective

(Castegnaro

etal.,

1997)

N/D

Absence

of

fluorescentby-

products

(Castegnaro

etal.,1997)

Noresidual

mutagenicity

(Castegnaro

etal.,

1997)

AP

PE

ND

IX.

(Contin

ued)

Pharm

aceutical

(CAS#)

Class

Applied

process

Degreeofdegradation

Kinetic

parameters

Interm

ediates/

by-products

Biodegradabilityand

toxicity

406 K. Ikehata et al. December 2006

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uly

2007

Epirubicin

(56420-45-2)

Cytostaticdrug,

anthracycline

Fenton

Complete

conversionof

2g/L

epirubicin

(Castegnaro

etal.,

1997),H

2O

2alonewas

less

effective

(Castegnaro

etal.,

1997)

N/D

Absence

of

fluorescentby-

products

(Castegnaro

etal.,1997)

Noresidual

mutagenicity

(Castegnaro

etal.,

1997)

Idarubicin

(58957-92-9)

Cytostaticdrug,

anthracycline

Fenton

Complete

conversionof

0.5

g/L

idarubicin

(Castegnaro

etal.,

1997),H

2O

2alonewas

less

effective

(Castegnaro

etal.,

1997)

N/D

Absence

of

fluorescentby-

products

(Castegnaro

etal.,1997)

Noresidual

mutagenicity

(Castegnaro

etal.,

1997)

Pirarubicin

(72496-41-4)

Cytostaticdrug,

anthracycline

Fenton

Complete

conversionof

1g/L

pirarubicin

(Castegnaro

etal.,

1997),H

2O

2alonewas

less

effective

(Castegnaro

etal.,

1997)

N/D

Absence

of

fluorescentby-

products

(Castegnaro

etal.,1997)

Noresidual

mutagenicity

(Castegnaro

etal.,

1997)

Azathioprine

(446-86-6)

Cytostaticdrug,

anti-m

etabolite

Ozonation

Complete

conversionof

499mg/L

azathioprine

atpH

3in

45min

(Rey

etal.,1999)

N/D

N/D

Noresidual

mutagenicity(R

eyet

al.,1999)

Cytarabine

(147-94-4)

Cytostaticdrug,

anti-m

etabolite

Ozonation

Complete

conversionof

401mg/L

cytarabine

atpH

3and7in

60-75

min

(Rey

etal.,1999)

N/D

N/D

Noresidual

mutagenicity(R

eyet

al.,1999)

5-Fluorouracil

(51-21-8)

Cytostaticdrug,

anti-m

etabolite

Ozonation

Complete

conversionof

269mg/L

5-

fluorouracilatpH

3and7in

45min

(Rey

etal.,1999)

N/D

N/D

Noresidual

mutagenicity(R

eyet

al.,1999)

Methotrexate

(59-05-2)

Cytostaticdrug,

anti-m

etabolite

Ozonation

Complete

conversionof

491mg/L

methotrexate

atpH

3and7in

60-75min

(Rey

etal.,1999)

N/D

N/D

Noresidual

mutagenicity(R

eyet

al.,1999)

(Contin

ued)

Ozonation and Advanced Oxidation of Pharmaceuticals December 2006 407

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2007

Cyclophosphamide

(50-18-0)

Cytostaticdrug,

alkylating

agent

Fenton

Complete

conversionof

4g/L

cyclophosphamide

(Hanselet

al.,1997)

N/D

N/D

Mutagenic

by-products

form

edin

thepresence

of5%

dextrose

(Hanselet

al.,1997)

Ifosfamid

(3778-73-2)

Cytostaticdrug,

alkylating

agent

Fenton

Complete

conversionof

27g/L

ifosfamid

(Hanselet

al.,1997)

N/D

N/D

Noresidual

mutagenicity(H

ansel

etal.,1997)

Melphalan

(148-82-3)

Cytostaticdrug,

alkylating

agent

Fenton

Complete

conversionof

2g/L

melphalan

(Hanselet

al.,1997)

N/D

N/D

Noresidual

mutagenicity(H

ansel

etal.,1997)

Cim

etidine

(51481-61-9)

HistamineH

2-

receptor

antagonist

Fenton

N/D

k�O

H=

6.5

·109

M�1�s�

1(Latchet

al.,

2003)

Fourdegradation

products

identified

(Zbaidaet

al.,

1986)

N/D

Ranitidine

(66357-35-5)

HistamineH

2-

receptor

antagonist

Fenton

N/D

k�O

H=

1.5

·1010

M�1�s�

1(Latchet

al.,

2003)

N/D

N/D

TiO

2/hn

Complete

conversionof

50mg/L

ranitidinein

1h,60%

TOC

removalin

5h

(Addamoet

al.,2005)

Kinetic

model

presented

(Addamoet

al.,2005)

N/D

N/D

17b-Estradiol

(50-28-2)

Horm

one,

natural

estrogen

Ozonation

Complete

conversionof

0.02mg

/L–1.4

mg/L

of

17b-estradiol(O

nda

etal.,2002;Alum

etal.,2004;Kim

etal.,

2004;Huber

etal.,

2005)

N/D

Degradation

pathway

proposed(H

uber

etal.,2004)

Estrogenicityremoved

(Ondaet

al.,2002;

Alum

etal.,2004;Kim

etal.,2004;Liu

etal.,

2005)

O3/H

2O

2Complete

conversionof

7-36ng/L

17b-estradiol(O

nda

etal.,2002)

N/D

N/D

Estrogenicityremoved

(Shishidaet

al.,2000;

Ondaet

al.,2002)

H2O

2/U

VComplete

conversionof

17b-estradiol

(unknown

concentration)

(Rosenfeldtand

Linden,2004)

k�O

H=

1.41

·1010

M�1�s�

1,akinetic

model

presented

(Rosenfeldtand

Linden,2004)

N/D

N/D

AP

PE

ND

IX.

(Contin

ued)

Pharm

aceutical

(CAS#)

Class

Applied

process

Degreeofdegradation

Kinetic

parameters

Interm

ediates/

by-products

Biodegradabilityand

toxicity

408 K. Ikehata et al. December 2006

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25 J

uly

2007

Photo-Fenton-

like

>85%

conversionof

5mg/L

17b-estradiol

atpH

3.0,

interm

ediate

reactivity

amongfourestrogens

(Fenget

al.,2005)

N/D

N/D

N/D

TiO

2/hn

Complete

conversion

andmineralizationof

0.272mg/L

17b-estradiolin

4h(O

hkoet

al.,2002)

N/D

Three

interm

ediates

(Ohkoet

al.,

2002)

Estrogenicityremoved

(Ohkoet

al.,2002)

Estrone(53-16-7)

Horm

one,

natural

estrogen

Ozonation

Complete

conversionof

0.015mg

/Lestrone

(Ternes

etal.,2003),

incomplete

(>90%)

conversionof0.5

mg/L

estrone(H

uber

etal.,

2005)

N/D

Degradation

pathway

proposed(H

uber

etal.,2004)

N/D

Photo-Fenton-

like

Complete

conversion

and15%

mineralizationof

5mg/L

estronein

160

min

atpH

3,least

degradable

among

fourestrogens(Feng

etal.,2005)

Anapparentkinetic

equationreported

(Fenget

al.,2005)

Six

unidentified

interm

ediates

(Fenget

al.,

2005)

N/D

17a-Ethinylestradiol

(57-63-6)

Synthetic

estrogen,

contraceptive

Ozonation

Complete

conversionof

0.15mg/L

17a-

ethinylestradiolatpH

8and10� C

(Huber

etal.,2003)

kO3=

7·109M�1�s�

1

atpH

10(H

uber

etal.,

2003)

Fiveinterm

ediates

(Huber

etal.,

2004)

Estrogenicityremoved

(Alum

etal.,2004;

Huber

etal.,2004;Liu

etal.,2005)

H2O

2/U

VComplete

conversionof

17a-ethinylestradiol

(unknown

concentration)

(Rosenfeldtand

Linden,2004)

k�O

H=

1.08

·1010

M�1�s�

1(R

osenfeldt

andLinden,2004),9.8

·109M�1�s�

1(H

uber

etal.,2003),akinetic

model

presented

(Rosenfeldtand

Linden,2004)

N/D

N/D

(Contin

ued)

Ozonation and Advanced Oxidation of Pharmaceuticals December 2006 409

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Photo-Fenton-

like

>85%

conversionof

5mg/L

17a-

ethinylestradiolatpH

3.0,interm

ediate

reactivityamongfour

estrogens(Fenget

al.,

2005)

N/D

N/D

N/D

Diethylstilbestrol

(56-53-1)

Synthetic

estrogen,

contraceptive

Photo-Fenton-

like

Complete

conversionof

5mg/L

diethylstilbestrolatpH

3.0,highestreactivity

amongfourestrogens

(Fenget

al.,2005)

N/D

N/D

N/D

Bezafibrate

(41859-67-0)

Lipid

regulator

Ozonation

Complete

conversionof

upto

181mg

/Lbezafibrate

atpH

8and10� C

(Huber

etal.,2003),

incomplete

conversion

atlower

ozonedoses

(Ternes

etal.,2002)

kO3=

590M�1�s�

1at

pH

5-10,20� C

(Huber

etal.,2003)

N/D

N/D

H2O

2/U

VN/D

k�O

H=

7.4

·109

M�1�s�

1atpH

7,25� C

(Huber

etal.,2003)

N/D

N/D

AP

PE

ND

IX.

(Contin

ued)

Pharm

aceutical

(CAS#)

Class

Applied

process

Degreeofdegradation

Kinetic

parameters

Interm

ediates/

by-products

Biodegradabilityand

toxicity

410 K. Ikehata et al. December 2006

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2007

Clofibricacid

(882-09-7)

Lipid

regulator

metabolite

Ozonation

Complete

conversionof

0.12mg

/L–322mg/L

clofibricacid

(Andreozziet

al.,

2003b;Ternes

etal.,

2003),49%

TOC

removalin

1hatpH

5.0

(Andreozziet

al.,

2003b),incomplete

conversionatlower

ozonedoses(Zwiener

andFrimmel,2000;

Ternes

etal.,2002;

Ternes

etal.,2003),

quantitative

dechlorination

(Andreozziet

al.,

2003b)

kO3>

20M�1�s�

1

(Huber

etal.,2005),

apparentrate

constants=

29.8

M�1�s�

1atpH

2–

2550M�1�s�

1atpH

6.5

(Andreozziet

al.,

2003b),akinetic

model

presented

(Andreozziet

al.,

2003b)

N/D

Algalandprotozoan

toxicityelim

inated(as

amixture)(A

ndreozzi

etal.,2004)

O3/H

2O

298%

conversionof2mg

/LclofibricacidatpH

7and10� C

in10min

(Zwiener

and

Frimmel,2000)

N/D

N/D

N/D

H2O

2/U

V>90%

conversionof

322mg/L

clofibric

acidatpH

5.0

and25

� C,12%

TOC

removal,>80%

dechlorination

(Andreozziet

al.,

2003b)

FP=

1.08

·10–2

mol�E

instein�1atpH

5.5

and254nm,

k�O

H=

2.38

·109

M�1�s�

1,akinetic

model

presented

(Andreozziet

al.,

2003b)

N/D

Algaltoxicity

elim

inated,protozoan

toxicityreduced(asa

mixture)(A

ndreozzi

etal.,2004)

Fenton

N/D

4.7

·109M�1�s�

1atpH

3.5

and22� C

(Packer

etal.,2003)

N/D

N/D

TiO

2/hn

Complete

conversionof

0.5

mg/L

clofibricacid

atpH

6.5

in4min

(DollandFrimmel,

2004),pilotscale

experim

ents

perform

ed(D

olland

Frimmel,2005a)

Pseudo-1st-order

kinetics(D

olland

Frimmel,2004)

Degradation

pathway

proposed(D

oll

andFrimmel,

2004)

Notoxicityreduction

(asamixture)

(Andreozziet

al.,

2004)

(Contin

ued)

Ozonation and Advanced Oxidation of Pharmaceuticals December 2006 411

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2007

Fenofibricacid

(42017-89-0)

Lipid

regulator

metabolite

Ozonation

Complete

conversionof

0.13mg

/Lfenofibric

acidatpH

7.2

(Ternes

etal.,2003)

N/D

N/D

N/D

Gem

fibrozil

(25812-30-0)

Lipid

regulator

Ozonation

Presence

ofunknown

amountofgem

fibrozil

wasreported

ina

sewagetreatm

ent

planteffluentthatwas

treatedbyozonation

(Huber

etal.,2005)

N/D

N/D

N/D

Diatrizoate

(737-31-5;

sodium

salt)

Triiodinated

X-raycontrast

medium

Ozonation

0%

to14%

conversion

of5.7

mg/L

diatrizoate

atpH

7.2

(Ternes

etal.,2003),no

conversionof5.0

mg/L

diatrizoate

atpH

7(H

uber

etal.,2005)

N/D

N/D

N/D

O3/H

2O

225%

conversionof5.7

mg/L

diatrizoate

atpH

7.2

(Ternes

etal.,

2003)

N/D

N/D

N/D

O3/U

V36%

conversionof5.7

mg/L

diatrizoate

atpH

7.2

(Ternes

etal.,

2003)

N/D

N/D

N/D

Iomeprol

(78649-41-9)

Triiodinated

X-raycontrast

medium

Ozonation

34%

to90%

conversion

of2.3-5

mg/L

iomeprol

atneutralpH

(Ternes

etal.,2003;Huber

etal.,2005)

N/D

N/D

N/D

O3/H

2O

285%

conversionof2.3

mg/L

iomeprolatpH

7.2

(Ternes

etal.,

2003)

N/D

N/D

N/D

O3/U

V88%

conversionof2.3

mg/L

iomeprolatpH

7.2

(Ternes

etal.,

2003)

N/D

N/D

N/D

AP

PE

ND

IX.

(Contin

ued)

Pharm

aceutical

(CAS#)

Class

Applied

process

Degreeofdegradation

Kinetic

parameters

Interm

ediates/

by-products

Biodegradabilityand

toxicity

412 K. Ikehata et al. December 2006

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2007

TiO

2/hn

68%

of4.1

mg/L

iomeprolin

3min

at

pH

6.5,40%

deiodination(D

olland

Frimmel,2004),pilot

scale

experim

ents

perform

ed(D

olland

Frimmel,2005a)

Pseudo-1st-order

kinetics(D

olland

Frimmel,2004)

Unidentified

organic

by-

products(D

oll

andFrimmel,

2004)

N/D

Iopamidol

(62883-00-5)

Triiodinated

X-raycontrast

medium

Ozonation

33%

to84%

conversion

of1.1-5

mg/L

iopamidolatneutral

pH

(Ternes

etal.,

2003;Huber

etal.,

2005)

N/D

N/D

N/D

O3/H

2O

280%

conversionof1.1

mg/L

iopamidolatpH

7.2

(Ternes

etal.,

2003)

N/D

N/D

N/D

O3/U

V90%

conversionof1.1

mg/L

iopamidolatpH

7.2

(Ternes

etal.,

2003)

N/D

N/D

N/D

Iopentol

(89797-00-2)

Triiodinated

X-raycontrast

medium

H2O

2/U

VComplete

AOX

removal

and80%

TOC

removalof785mg/L

iopentolatpH

6.6,

85%

ofiodine

recovered

aselem

ental

iodine(Spreheet

al.,

2001)

N/D

N/D

N/D

Iopromide

(73334-07-3)

Triiodinated

X-raycontrast

medium

Ozonation

42%

to91%

conversion

of5–5.2

mg/L

iopromideatneutral

pH

(Ternes

etal.,

2003;Huber

etal.,

2005)

kO3<

0.8

M�1�s�

1at

pH

5to

10and20� C

(Huber

etal.,2003)

N/D

N/D

O3/H

2O

289%

conversionof5.2

mg/L

iopromideatpH

7.2

(Ternes

etal.,

2003)

k�O

H=

3.3

·109

M�1�s�

1atpH

7and

25� C

(Huber

etal.,

2003)

N/D

N/D

(Contin

ued)

Ozonation and Advanced Oxidation of Pharmaceuticals December 2006 413

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2007

O3/U

V90%

conversionof5.2

mg/L

iopromideatpH

7.2

(Ternes

etal.,

2003)

N/D

N/D

N/D

H2O

2/U

VComplete

AOX

removal,>90%

TOC

removalof769mg/L

iopromideatpH

6.6

(Spreheet

al.,2001)

N/D

N/D

N/D

TiO

2/hn

Sim

ilardegradabilityto

iomeprol(D

olland

Frimmel,2004)

N/D

N/D

N/D

Note:N/D

=notdetermined,TOC=

totalorganic

carbon,BOD

5=

5-daybiochem

icaloxygen

dem

and,COD=

chem

icaloxygen

dem

and,

DOC=

dissolved

organicmatter,AOX=

absorbableorganichalogens,kO3=

second-order

rate

constantformolecularozonereaction,k�O

H=

second-order

rate

constantforhydroxylradicalreaction,FP=

quantum

yield

ofdirectphotolysis.

AP

PE

ND

IX.

(Contin

ued)

Pharm

aceutical

(CAS#)

Class

Applied

process

Degreeofdegradation

Kinetic

parameters

Interm

ediates/

by-products

Biodegradabilityand

toxicity

414 K. Ikehata et al. December 2006