evaluation of advanced oxidation processes for water and

lable at ScienceDirect

Water Research 139 (2018) 118e131

Contents lists avai

Water Research

journal homepage: www.elsevier .com/locate/watres

Review

Evaluation of advanced oxidation processes for water and wastewatertreatment e A critical review

David B. Miklos a, Christian Remy b, Martin Jekel c, Karl G. Linden d, J€org E. Drewes a,Uwe Hübner a, *

a Chair of Urban Water Systems Engineering, Technical University of Munich, Am Coulombwall 3, D-85748 Garching, Germanyb Kompetenzzentrum Wasser Berlin gGmbH, Cicerostrasse 24, D-10709 Berlin, Germanyc Technische Universit€at Berlin, Chair of Water Quality Control, KF4, Str. des 17. Juni 135, D-10623, Berlin, Germanyd Department of Civil, Environmental, and Architectural Engineering, University of Colorado Boulder, UCB 607, Boulder, CO 80303, USA

a r t i c l e i n f o

Article history:Received 19 February 2018Received in revised form13 March 2018Accepted 15 March 2018Available online 22 March 2018

Keywords:Advanced oxidation processTrace organic chemicalsOxidation by-productsOH-Radical exposureElectrical energy per order (EEO)

* Corresponding author.E-mail addresses: [email protected] (D.B. Miklos),

wasser.de (C. Remy), [email protected] (M.edu (K.G. Linden), [email protected] (J.E. Drewes), u.hu

https://doi.org/10.1016/j.watres.2018.03.0420043-1354/© 2018 Elsevier Ltd. All rights reserved.

a b s t r a c t

This study provides an overview of established processes as well as recent progress in emerging tech-nologies for advanced oxidation processes (AOPs). In addition to a discussion of major reaction mech-anisms and formation of by-products, data on energy efficiency were collected in an extensive analysis ofstudies reported in the peer-reviewed literature enabling a critical comparison of various established andemerging AOPs based on electrical energy per order (EEO) values. Despite strong variations withinreviewed EEO values, significant differences could be observed between three groups of AOPs: (1) O3

(often considered as AOP-like process), O3/H2O2, O3/UV, UV/H2O2, UV/persulfate, UV/chlorine, andelectron beam represent median EEO values of <1 kWh/m3, while median energy consumption by (2)photo-Fenton, plasma, and electrolytic AOPs were significantly higher (EEO values in the range of 1e100kWh/m3). (3) UV-based photocatalysis, ultrasound, and microwave-based AOPs are characterized bymedian values of >100 kWh/m3 and were therefore considered as not (yet) energy efficient AOPs. Spe-cific evaluation of 147 data points for the UV/H2O2 process revealed strong effects of operational con-ditions on reported EEO values. Besides water type and quality, a major influence was observed forprocess capacity (lab-vs. pilot-vs. full-scale applications) and, in case of UV-based processes, of the lamptype. However, due to the contribution of other factors, correlation of EEO values with specific waterquality parameters such as UV absorbance and dissolved organic carbon were not substantial. Also,correlations between EEO and compound reactivity with OH-radicals were not significant (photolyticallyactive compounds were not considered). Based on these findings, recommendations regarding the use ofthe EEO concept, including the upscaling of laboratory results, were derived.

© 2018 Elsevier Ltd. All rights reserved.

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1192. Background regarding advanced oxidation processes for contaminant removal in water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 120

2.1. Ozone based AOPs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1212.1.1. Ozonation at elevated pH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1212.1.2. Peroxone-process (O3/H2O2) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1212.1.3. O3/catalysts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 121

2.2. UV-based AOPs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1212.2.1. UV/H2O2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1222.2.2. UV/O3 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 122

Christian.Remy@kompetenz-Jekel), [email protected]@tum.de (U. Hübner).

mailto:[email protected]








http://crossmark.crossref.org/dialog/?doi=10.1016/j.watres.2018.03.042&domain=pdf

www.sciencedirect.com/science/journal/00431354

www.elsevier.com/locate/watres

https://doi.org/10.1016/j.watres.2018.03.042



D.B. Miklos et al. / Water Research 139 (2018) 118e131 119

2.2.3. UV/Cl2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1222.3. Electrochemical AOPs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1222.4. Catalytic AOPs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 123

2.4.1. Fenton process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1232.4.2. Photocatalytic AOPs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 123

2.5. Physical AOPs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1232.5.1. Electrohydraulic discharge (Plasma) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1232.5.2. Ultrasound . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1232.5.3. Microwave . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1242.5.4. Electron beam . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 124

3. Oxidation by-products . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1243.1. Reactions with inorganic compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1243.2. Reactions with organic compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 124

4. Comparison of advanced oxidation processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1254.1. Comparative screening of EEO values for different AOPs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1254.2. Principal influences on EEO-values shown at the UV/H2O2 process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 127

4.2.1. Influence of process capacity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1274.2.2. Influence of compound reactivity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1274.2.3. Influence of water quality . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1274.2.4. Influence of lamp type . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 128

5. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 128Supplementary data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 128References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 128

Abbreviations

AOP Advanced oxidation processesAOX Adsorbable organic halidesBDD Boron-doped diamondcAOP Catalytic advanced oxidation processesCOD Chemical oxygen demandDOC Dissolved organic carbonDOM Dissolved organic mattereAOP Electrochemical advanced oxidation processesEEO Electrical energy per orderHAA Haloacetic acidsHAN HaloacetonitrilesIUPAC International Union of Pure and Applied Chemistry

LED Light emitting diodeLP Low-pressureMP Medium-pressureNDMA N-NitrosodimethylamineNTU Nephelometric turbidity unitOBPs Oxidation by-productspAOP Physical advanced oxidation processesPDS PeroxydisulfatePMS PeroxymonosulfateTHM TrihalomethanesTOrCs Trace organic chemicalsTOX Total organic halidesUS UltrasoundUVA UV absorbance

1. Introduction

In recent years, trace organic chemicals (TOrC) such as phar-maceuticals, consumer products, and industrial chemicals havebeen detected in the aquatic environment (Huerta-Fontela et al.,2010). Besides urban and agricultural run-offs, wastewater treat-ment plant effluents are considered to be the most significant TOrCemitters (Lim, 2008; Gros et al., 2010; Luo et al., 2014). TOrCsremain in wastewater treatment plant effluents being dischargedinto surface waters, since conventional physical and biologicalwastewater treatment can only partially remove these substances(Lim, 2008; Zhang et al., 2008; Luo et al., 2014).

The application of advanced oxidation processes (AOPs) pro-vides a viable and effective attenuation option due to the oxidationof a wide range of TOrCs (Comninellis et al., 2008; Klavarioti et al.,2009; Yang et al., 2014; Giannakis et al., 2015; Stefan, 2018). Ac-cording to the definition of Bolton et al. (1996) and Bolton et al.(2001), AOPs are based on the in situ generation of strong oxi-dants for the oxidation of organic compounds. This includes pro-cesses based on OH-radicals (�OH), which constitute the majority ofavailable AOPs, but also processes based on other oxidizing species

favoring sulfate or chlorine radicals. There are various differentprocess technologies which have been investigated for use as AOPs.Several AOPs, especially those involving ozonation and UV irradi-ation are already well established and operated at full-scale indrinking water treatment and water reuse facilities. However, newstudies of numerous emerging AOPs for water treatment (i.e.,electrochemical AOP, plasma, electron beam, ultrasound or micro-wave based AOPs) are constantly being reported by various re-searchers (Stefan, 2018). The huge amount of different studies andan increasing number of proposed technologies and process com-binations pose an enormous challenge for a critical assessment ofAOPs concerning their operational costs (i.e., energy consumption,chemical input), sustainability (i.e., resource use, carbon footprint),and general feasibility (e.g., physical footprint and oxidation by-product formation) to enable comparison of their efficiency withother AOPs and alternative treatment processes.

To address this issue, Bolton and coworkers developed figures ofmerit for the comparison of advanced oxidation processes (Boltonet al., 2001). These are based on electrical energy consumptionwhich often represents a major fraction of the AOP operating costs.For low contaminant concentrations (typically< 100mg/L), the

D.B. Miklos et al. / Water Research 139 (2018) 118e131120

kinetics of destruction of organic contaminants by AOPs can oftenbe described phenomenologically by simple pseudo first-order rateexpressions. Thus, the oxidant or energy dosage scales with thevolume and treatment goals (i.e. orders of magnitude of reductionper unit volume). Consequently, the figure of merit for electrical-driven AOPs is defined as EEO (electrical energy per order):

“Electrical energy per order is the electrical energy in kWh requiredto degrade a contaminant C by one order of magnitude in 1 m3 ofcontaminated water” (Bolton et al., 1996).

This figure of merit has been accepted by the InternationalUnion of Pure and Applied Chemistry (IUPAC) in 2001 (Bolton et al.,2001) and numerous EEO values have been reported since then inliterature for various oxidation processes and applications. Giving adirect link to the electrical efficiency of the AOP, this approach al-lows not only for a simple comparison of different AOP technolo-gies, but also provides the requisite data for scale-up and economicas well as sustainability analyses for comparison with conventionaltreatment technologies (e.g., activated carbon adsorption, airstripping).

In aqueous systems, oxidation of a specific compound C followsa second-order reaction, where the relative residual concentrationis a function of compound specific rate constant kOH and the �OHexposure. Accordingly, �OH exposure can be determined fromexperimental data using equation (1).

Zð�OHÞdt ¼

ln�CC0

��k�OH; S

(1)

The �OH exposure is controlled by the radical formation effi-ciency of the respective process as well as competing reactions withother constituents in the water called radical scavenging. Major

Fig. 1. Broad overview and classification of different AOPs. Individual processes are marked aat lab-scale (black).

radical scavengers are carbonate, bicarbonate, nitrite, and organicmatter indicating a strong dependency of compound removal andthus EEO values on the water matrix. Besides radical scavenging, thewater matrix might also directly affect the in situ generation ofradicals in several processes, e.g. by reducing UV-transmittance orreactions with ozone in ozone-based AOPs. For these reasons, theapplication of EEO values for a comparison of experimental resultsfrom different water matrices is not recommended and compara-tive studies to evaluate efficiency of different AOPs in a definedwater matrix are needed (Bolton et al., 2001). To date, only fewstudies directly comparing different AOPs are available (Boltonet al., 1998; Müller et al., 2001; Alaton et al., 2002; Katsoyianniset al., 2011; Ure~na de Vivanco et al., 2013; Lutterbeck et al., 2015;Fast et al., 2017) and they are mostly limited to a few establishedprocesses. To the best of our knowledge, such a comprehensivecomparison across different AOPs has not yet been conducted.

This article provides a critical review of different established andemerging AOPs based on data compiled during an extensive liter-ature study. An initial comparative assessment is conducted basedon EEO-values reported in the peer-reviewed literature for differentAOPs. Influencing aspects, such as reaction rate constants of targetsubstances, water matrix, process capacity or system parametersare considered and critically evaluated. As a result, recommenda-tions for the use of the EEO-concept in future studies are presented.In addition, this article also provides an assessment on by-productformation in different AOPs based on reaction mechanisms ofdifferent oxidants.

2. Background regarding advanced oxidation processes forcontaminant removal in water

Technologies for AOPs involve widely different methods ofactivation as well as oxidant generation and can potentially utilize a

s established at full-scale (white), investigated at lab- and pilot-scale (grey), and tested


number of different mechanisms for organic destruction. An over-view of different established and emerging AOPs is given in Fig. 1,categorized into ozone-based, UV-based, electrochemical (eAOP),catalytic (cAOP), and physical (pAOP) AOPs. However, it is note-worthy that this classification scheme should not be viewed asstrict since several processes involve different technologies andthus could be assigned to various categories. The different pro-cesses summarized in Fig. 1 represent processes of very differentdegrees of implementation from well-established AOPs to pro-cesses only tested at laboratory scale yet.

All AOPs comprise of two steps, the in situ formation of reactiveoxidative species and the reaction of oxidants with target con-taminants. Mechanisms of radical formation depend on processspecific parameters and can be affected by system design and waterquality. Besides radical scavenging also other parameters (e.g.,radical mass transfer in surface based AOPs, hydrodynamics) playan important role for efficiency of contaminant destruction. In thefollowing sections, the current status of implementation isreviewed, major mechanisms and principles of radical generationare illustrated, and constraints of different AOPs are briefly dis-cussed. More comprehensive overview on system design, reactionprinciples and kinetics can be found in various book publications onAOPs (Parsons, 2004; Collins and Bolton, 2016; Stefan, 2018).Mechanisms for the formation of oxidation by-products (OBPs) indifferent AOPs are discussed separately in section 3.

2.1. Ozone based AOPs

Ozone has long been used as an oxidant and disinfectant inwater treatment. As an oxidant, ozone is very selective and attacksprimarily electron-rich functional groups like double bonds,amines, and activated aromatic rings (e.g. phenol). Since its re-actions in real aqueous solutions often involve the formation of�OH, ozonation itself is often considered an AOP or AOP-like pro-cess. �OH can be formed from the reaction of ozone with hydroxideions (Mer�enyi et al., 2010a, 2010b). The initiation of this reaction,however, is quite slow with a second-order rate constant of70M�1s�1.

In addition, radicals are formed as a side product from the re-action of ozone with organic matter (mainly phenol and aminefunctional groups) (Buffle and von Gunten, 2006). Especially duringozonation of secondary effluents these reactions are the majorcontributors to radical formation. Methods to actively initiate for-mation of radicals include the ozonation at elevated pH and thecombinations O3/H2O2 (also called peroxone-process), O3/UV, andO3/catalysts. The combination of ozonation and UV-irradiation willbe discussed as a UV-based AOP in section 2.2.

2.1.1. Ozonation at elevated pHOzonation at elevated pH is considered as an AOP if �OH gen-

eration is intentionally favored (Elovitz and von Gunten, 1999;Buffle et al., 2006). The pH of treated water influences directozonation efficiency since dissociated target organic compoundsmight have significantly different kO3 values (Calderara et al., 2002).Furthermore, the abundance of hydroxide ions directly influencesthe �OH generation and therefore indirect ozonation. Especially ifthe water to be treated has a pH> 8, ozonation applied as an AOPmight be a promising process, if the precipitation of calcium car-bonate is not of concern.

2.1.2. Peroxone-process (O3/H2O2)In the peroxone process, ozone reacts with the peroxide anion

(HO2�) to form �OH precursors, which are subsequently reacting to

�OH. For a detailed mechanistic description of the peroxone processsee Mer�enyi et al. (2010a). Residual H2O2 might have to be

destroyed before discharging the treated water to the receivingaqueous environment. The optimum molar ratio for the peroxoneprocess is H2O2/O3¼ 0.5mol/mol (Katsoyiannis et al., 2011;Pisarenko et al., 2012). Typical ozone doses in the peroxone processare 1e20mg/L. Peroxide can also be formed from reactions ofozone with the water matrix but its contribution to overall �OHformation during wastewater ozonation is not significant (N€otheet al., 2009). O3/H2O2 is a well-established process in drinkingwater treatment and water reuse applications (e.g. Windhoek,Namibia). However, recent studies have shown that benefits for itsapplication in wastewater are limited due to high competition re-actions and already efficient radical formation with ozone alone(Hübner et al., 2015). However, it might still be a valuable treatmentoption to minimize bromate formation during ozonation as dis-cussed in section 3.

2.1.3. O3/catalystsCatalytic ozonation is distinguished between homogeneous and

heterogeneous catalytic ozonation, depending on the water solu-bility of the catalyst. Homogeneous catalytic ozonation can bedescribed as a three-step catalytic cycle as approached by Pines andReckhow (2002) using Co(II) as a catalyst and oxalic acid: (1) for-mation of Co(II)-oxalate complex, (2) oxidation by ozone to Co(III)-oxalate complex, and (3) decomposition of Co(III)-oxalate complexforming an oxalate radical and Co(II). Heterogeneous catalyticozonation mechanisms are mediated by metal oxides (e.g., TiO2,Al2O3, MnO2) and result in more complex reaction paths based onmultiple-phase transport mechanisms and respective reactions asdescribed in detail by Beltr�an (2004).

Both homogeneous and heterogeneous catalytic ozonation haveshown their potential for water treatment at laboratory scalemainly based on lower ozone demand compared to ozonationalone (Bai et al., 2016; Wu et al., 2016a, 2016b; Xing et al., 2016).However, full-scale application is limited due to catalyst recoveryand a lack of understanding of the catalytic ozonation mechanisms(Nawrocki and Kasprzyk-Hordern, 2010). Some studies report theuse of activated carbon as a catalyst in catalytic ozonation (e.g.Kaptijn, 1997). However, �OH production in this process is based onthe reaction of ozone with pyrrol groups present on the activatedcarbon surface indicating that it acts rather as a radical promoterthan a catalyst, which needs to be continuously renewed tomaintain efficient radical generation (S�anchez-Polo et al., 2005).

2.2. UV-based AOPs

UV-based AOPs comprise processes based on UV-irradiation(mostly UV-C) and the combination of UV-light with differentradical promoters. UV-fluences applied for advanced oxidation areusually >200mJ/cm2 and therefore exceed UV-dose requirementsfor 4-log inactivation of most pathogens including UV-resistantorganisms (e.g. adenovirus) (EPA, 2006). UV-irradiation sourcesusually consist of either low- (LP) or medium-pressure (MP) mer-cury lamps with mono- or polychromatic emission spectra,respectively. Recently, UV-light emitting diode (LED) light sourceswith specific wavelength distributions have been investigated andsummarized for disinfection purposes (Song et al., 2016). Theprincipal advantages of LEDs compared to conventional mediumand low-pressure lamps are the elimination of mercury, uniquepeak emission wavelengths, compact size and therefore flexibleapplication design as well as a short start-up phase. However,despite the prediction of future UV-LED wall plug efficiencies ofabout 75% in 2020 (Autin et al., 2013), current diodes emit UV ra-diation at efficiencies of <10% (Chen et al., 2017). This results in EEOvalues for LED systems that are not yet competitive with conven-tional UV-systems (Wang et al., 2017) and are therefore not


considered in this study.The most frequently applied UV-based AOP is the combination

with H2O2. Other radical promoters such as persulfate (to formsulfate radicals) and chlorine (hydroxyl radicals and radical chlo-rine species) are also being investigated. Besides established oxi-dants, Keen et al. (2012) investigated the applicability of nitrate incombination with MP-lamps as an alternative UV-based AOP.However, to the best of our knowledge, no EEO values are availablefor this process.

2.2.1. UV/H2O2

The combination of UV-irradiation and H2O2 leads to the photo-lytic cleavage of H2O2 into two �OH. However, the molar absorptioncoefficient of H2O2 is relatively low with ε¼ 18.6M�1cm�1 atl¼ 254 nm resulting in a H2O2 turnover of <10%. If LP UV lamps areused, high concentrations of H2O2 are required to generate sufficient�OH ([H2O2]¼ 5e20mg/L) leading to the necessity of removingresidual H2O2 in a subsequent step. Applied H2O2-doses are mainlyset based on economic aspects. However, at higher concentrationsalso scavenging of �OHwithH2O2 (kOH,H2O2¼ 2.7 $107M�1s�1)mightaffect the radical yield (Buxton et al., 1988).

UV/H2O2 for TOrC removal has been examined widelythroughout peer-reviewed journal articles at lab-scale (Wols andHofman-Caris, 2012; Wols et al., 2013; Keen et al., 2016) for waterqualities ranging from ultrapure water to landfill leachate (Xiaoet al., 2016; Ghazi et al., 2014). First full-scale applications arealready established for potable water reuse (Audenaert et al., 2011)and surface water treatment applications (Kruithof et al., 2007).UV/H2O2 is not established for advanced wastewater treatmentmainly because of low UV-transmittance and high scavenging ca-pacity of secondary or tertiary treated wastewater effluents but isused in some potable reuse treatment trains employing integratedmembrane systems (ultrafiltration/reverse osmosis) (Drewes andKhan, 2015) based on its negligible OBP formation potential asdiscussed in section 3.

2.2.2. UV/O3

In the UV/O3 process, UV irradiation (l< 300 nm) results in acleavage of dissolved ozone, followed by a fast reaction of atomicoxygen with water to form a thermally excited H2O2. Subsequently,the excited peroxide decomposes into two �OH (von Sonntag,2008). Ozone has a molar extinction coefficient ofε¼ 3300M�1cm�1 at l¼ 254 nm, which is significantly higher thanthat of H2O2 at this particular wavelength. However, due to cagerecombinations only a small proportion of generated H2O2 de-composes to �OH resulting in a free �OH quantum yield of only 0.1(Reisz et al., 2003). Furthermore, both UV lamps and ozonegenerator need large amounts of electrical energy, resulting inrelatively high energy demands for the combination of UV andozone. Direct oxidation by the combination of ozonation andphotolysis covers a wide range of TOrC reactivity and leads to themain advantage of this process. However, low energy efficiency ofradical generation might explain that to the best of our knowledge,no published data on full-scale UV/O3 application are available.

2.2.3. UV/SO4�-

An interesting alternative to �OH based AOPs is UV/SO4�- which

generates primarily sulfate radicals (SO4�-) for the oxidation of

organic contaminants in water (Lutze, 2013; Ao and Liu, 2016;Wacławek et al., 2017; Ike et al., 2018). Sulfate radicals have a strongoxidizing power and are more selective oxidants than �OH (Lutzeet al., 2015).

Peroxydisulfate (PDS, S2O82�) is homolytically cleaved by UV-C

activation. The quantum yield of S2O82� is larger than H2O2 (1.4

compared to 1.0) and molar absorption for S2O82� is slightly higher

as well (22M�1cm�1 and 18.6M�1cm�1, respectively) resulting in ahigher generation of radicals using PDS as oxidizing agent (Legriniet al., 1993; Lutze, 2013; Xiao et al., 2016). Peroxymonosulfate (PMS,HSO5

�) is activated by UV radiation into a SO4�- and a �OH with a

quantum yield of 0.52 at pH 7 (Guan et al., 2011). Several studieshave investigated the mechanisms and application of UV/PMS(Antoniou et al., 2010; Khan et al., 2014; Mahdi-Ahmed and Chiron,2014). However, based on its lower quantum yield, high commer-cial pricing and low availability of EEO values it is not considered inthis study (Wacławek et al., 2017).

Recent research has shown the advantages of UV/SO4�-

compared to UV/H2O2 in lab-scale experiments (Khan et al., 2014;Zhang et al., 2015; Xiao et al., 2016). However, based on more se-lective reactivity of sulfate radicals, results reveal a higher sensi-tivity to water matrix changes and DOM composition compared toUV/H2O2 (Ahn et al., 2017). Depending on the respective targetcompound and water matrix, SO4�

- based AOPs can be a consider-able alternative to �OH-based processes. However, UV/PDS yields inhigher formation potential of OBPs in comparison to UV/H2O2 (seesection 3).

2.2.4. UV/Cl2UV/Cl2 is another promising AOP, where UV-activated chlorine

forms radical species, i.e. Cl� and �Cl2� and �OH which then oxidizetarget compounds (Watts and Linden, 2007). Cl� is a more selectiveoxidant than �OH, since it reacts favorably with electron-rich con-taminants (Fang et al., 2014). The two oxidants mainly used arehypochlorite and chlorine dioxide (Jin et al., 2011; Sichel et al.,2011; Fang et al., 2014; Wang et al., 2016). However, regardinghypochlorite, pH dependency of HOCl/OCl� speciation needs to beconsidered since it influences the molar absorption coefficientsignificantly. UV/Cl2 is especially favorable for waters with lowerpH values such as reverse osmosis permeate (Watts et al., 2007).Research has mainly been conducted on lab-scale systemsdegrading organic indicator compounds (Jin et al., 2011; Sichelet al., 2011; Fang et al., 2014; Wang et al., 2016). A first full-scaleapplication for indirect potable reuse recently started operationat the Los Angeles Terminal IslandWater Reclamation Plant (Xylem,2015). However, Cl� based reactions involve the formation ofoxidative chlorine species (e.g., ClO�, OCl�), which might beoxidized by �OH to chlorate, perchlorate and halogenated OBPs (seesection 3 for more details).

2.3. Electrochemical AOPs

Electrochemical AOPs for water treatment applications wererecently reviewed in detail by Chaplin (2014). The major electrodetypes commonly used in this process are doped SnO2 (Zhuo et al.,2011), PbO2 (Bonfatti, 1999; Fernandes et al., 2014), RuO2 (Quanet al., 2013), boron-doped diamond (BDD) (Chaplin et al., 2013),and sub-stoichiometric and doped-TiO2 (Kesselman et al., 1997;Bejan et al., 2009). However, BDD-electrodes are the most appliedeAOP method due to their relatively low production costscompared to other electrodes and high stability of the diamondlayer under anodic polarization (Chaplin, 2014).

The electrochemical oxidative treatment of contaminated waterwith BDD electrodes can generate �OH directly via O2 evolutionfrom water oxidation (Tr€oster et al., 2004). As diamond is a non-conductor it is doped with boron to use it as an electrode mate-rial that is deposited onto a carrier material such as niobium,tantalum or silicon by chemical vapor deposition (Haenni et al.,1998). The radicals are generated without the addition of furtherchemicals. Therefore, BDD-electrode treatment attracts interest asan eco-friendly and efficient method for the removal of variouspollutants. However, since �OH generation occurs directly on the


electrode surface and reactivity range of �OH is limited to about1 mm (Kapałka et al., 2009), diffusive transport through theboundary layer at the electrode surface is the limiting factor of highoxidation efficiencies. For eAOP processes, hydrodynamic param-eters therefore have to be considered, as energy used to pumpwater, might account for the greatest share of energy consumptionin this process. This applies especially if low current densities areused to achieve higher �OH formation efficiency prolonging overalltreatment duration and pumping energy requirements.

Apart from the oxidation of TOrCs in water treatment, BDD-electrodes are investigated for disinfection purposes as well as forthe removal of chemical oxygen demand (COD) (Rajab et al., 2013,2015). Besides the generation of �OH, secondary oxidants, whichenhance elimination reactions and disinfection in the bulk solution,can be produced (Rajab et al., 2015). A limiting factor for theapplicability of BDD is unintentional formation of halogenatedOBPs as discussed in section 3 (von Gunten, 2003b; Bergmann andRollin, 2007; Bergmann et al., 2011). Nevertheless, several full-scaleeAOP systems for COD removal are already applied (Woisetschl€ageret al., 2015).

2.4. Catalytic AOPs

2.4.1. Fenton processThe combination of ferrous iron (Fe(II)) and H2O2 at acidic

conditions results in �OH formation (Fenton reaction). Iron acts as acatalyst with maximum catalytic activity at pH¼ 3, particularly dueto the precipitation of ferric oxyhydroxide at higher pH value(Wadley and Waite, 2004). Excess addition of H2O2 leads to thereduction of Fe(III) to Fe(II) (Safarzadeh-Amiri et al., 1996). Bysubstitution of iron oxides by other transition metals, enhancedTOrC removal performance can be achieved (Jiang et al., 2010;Rahim Pouran et al., 2014; Piscitelli et al., 2015). To prevent ironprecipitation, the Fenton process is restricted to acidic conditions.Therefore, alternative iron-free Fenton-like processes have recentlybeen investigated as summarized by Bokare and Choi (2014). Mainadvantages of the Fenton process are operation at low-costs(S�anchez Perez et al., 2013) and possibility of easy magnetic sepa-ration of residual iron. The Fenton process is therefore establishedin several industrial full-scale applications (e.g. Bae et al. (2015)).

2.4.2. Photocatalytic AOPsThe use of photo-active catalysts for oxidation processes in

water treatment has been investigated intensively over the lastdecades (Blake, 2001; Dong et al., 2015; Vallejo et al., 2015).Although there are numerous catalysts with photocatalytic prop-erties (i.e., TiO2, WO3 or ZnO), research has mainly concentrated ontwo types of reactions based on the solubility of the catalyst:

homogeneous photo-Fenton processes:

FeðOHÞ2þ þ hn/Fe2þ þ ,OH (2)

heterogeneous photocatalysis based on semiconductors (TiO2)(Simonsen et al., 2010):

TiO2 þ hn/�e� þ hþ

�(3)

hþ þ OH�ad / ,OHad (4)

UV and visible light (l¼ 180e400 nm) accelerate the Fentonprocess by photoreduction of Fe(III), however, the quantum yieldfor this reaction is relatively low (Wadley and Waite, 2004). Hence,

it is directly coupled with the Fenton process. Photo-Fenton pro-cesses with an organic ligand (e.g. ferrioxalate) have a higherquantum yield and thus a higher efficiency due to the high UVabsorption of Fe(III)-polycarboxylates. Additionally, the ferrioxalatecomplex can absorb radiation up to a wavelength of l< 550 nm,making it suitable for solar-driven AOPs (Hislop and Bolton, 1999).A recent review of photo-Fenton applications for wastewatertreatment is given by Rahim Pouran et al. (2015).

In TiO2-based photocatalysis, a semiconductor material is irra-diated by UV light (l< 400 nm). It is usually investigated as sus-pended colloidal particles or immobilized on different substrates. Ifphotons with sufficient energy hit the photocatalyst surface, anelectron is excited to the conduction band, leaving a positivelycharged hole (hþ) in the valence band (eq. (3)). These species cancause oxidative or reductive transformations of water constituents,either directly on the semiconductor surface or via radical reactions(eq. (4)). A sufficient amount of dissolved oxygen is necessary forthe latter reactions. The combination of oxidation and reductionmechanisms is specific for photocatalysis, whereas other AOPs arebased only on �OH reactions. Unfortunately, the quantum yield ofTiO2 photocatalysis for oxidation and reduction of contaminants isusually very low (f¼ 0.04) due to the fast recombination ofelectron-hole pairs (Sun and Bolton, 1996). Addition of an electrondonor (e.g. citric acid) may lead to the “filling” of positive holes andincreased reduction rates from the negative electrons in the con-duction band (Vohra and Davis, 2000; Oliveira et al., 2015).

Advantages of TiO2 photocatalysis for TOrC removal include lowcosts of the catalyst itself and easy commercial availability invarious crystalline forms and particle characteristics. Furthermore,the catalyst is non-toxic and photochemically stable. The limitationof heterogeneous photo-catalysis application at full-scale is mainlybased on two factors: (1) separation of colloidal catalyst from thewater suspension after treatment and (2) mass transfer limitationsto the surface of the immobilized catalyst on a substrate (Qu et al.,2013). Despite strong research efforts in the field of photocatalysis,the process is rarely applied in industrial or municipal watertreatment facilities because of the low quantum yield for �OHradical production.

2.5. Physical AOPs

2.5.1. Electrohydraulic discharge (Plasma)Liquid-phase electrical discharge reactors have been investi-

gated as AOPs inwater treatment (Locke et al., 2006; Hijosa-Valseroet al., 2014). Strong electric fields applied within the water (elec-trohydraulic discharge) or between water and gas phase(nonthermal plasma) initiate both chemical and physical processes.Beside the direct oxidation of contaminants in the water, variousoxidizing radicals or active species, UV radiation, and shock wavesare formed during the discharge, which can promote oxidation(Jiang et al., 2014).

2.5.2. UltrasoundSonication of water by ultrasound (US) (20e500 kHz) leads to

the formation and collapse of micro-bubbles from acoustical wave-induced compression and rarefaction. These bubbles implodeviolently after reaching a critical resonance size and generatetransient high temperatures (>5000 K), high pressures (>1000 bar)and highly reactive radicals. Destruction of water contaminantsoccurs by thermal decomposition and various radical reactions(Mason and P�etrier, 2004). Cavitation via ultrasound exhibits lowinterference from water matrix and less heat transfer compared toUV irradiation. A comprehensive review of sonochemical methodsis provided by Pang et al. (2011). Sonochemical processes haveproven to oxidize various aquatic contaminants in lab-scale


experiments (Mahamuni and Adewuyi, 2010). However, theapplication of ultrasound is highly energy intensive and results in avery low electrical efficiency of this AOP in comparison to othertechnologies (Goel et al., 2004; Mahamuni and Adewuyi, 2010).Therefore, the coupling of ultrasound with UV-irradiation (sono-photolysis), oxidants (O3, H2O2), or catalysts (TiO2) (sonocatalysis)or both (sonophotocatalysis) receives increased attention. Thesehybrid processes can yield additional advantages. However, majorimprovement of energy efficiency is often achieved due to thehigher efficiency of the coupled additional processes (e.g. UV/H2O2in US/UV/H2O2) (Mahamuni and Adewuyi, 2010).

2.5.3. MicrowaveThe application of highly energetic radiation in the microwave

range (300MHze300 GHz) has been investigated for the oxidationof water contaminants. Microwaves have been used in combinationwith oxidants (H2O2) or catalysts (TiO2, GAC) to assist in thedestruction of organic pollutants (Han et al., 2004; Zhihui et al.,2005; Bo et al., 2006). Microwaves can enhance reaction ratesand induce selective heating of the contaminants through internalmolecule vibration. Additionally, microwaves can generate UV ra-diation via an electrodeless discharge lamp for combined MW/UVreactors. Unfortunately, most of the applied microwave energy isconverted into heat. Beside the low electrical efficiency (EEO notreadily available in literature), cooling devices have to be employedto prevent treated water from overheating.

2.5.4. Electron beamThe utilisation of ionizing radiation from an electron beam

source (0.01e10MeV) for water treatment has been tested since the1980s. The accelerated electrons penetrate the water surface andresult in the formation of electronically excited species in thewater,including various ionic species and free radicals. The maximumpenetration depth of the accelerated electrons is directly propor-tional to the energy of the incident electrons (e.g. 7mm, reportedby Nickelsen (1994)). Therefore, water is irradiated in a thin film oras a sprayed aerosol. This process exhibits a high oxidizing powerand little interference by the water matrix and the electrical effi-ciency is within the feasibility range (EEO < 3 kWh/m3*order formost contaminants (Bolton et al., 1998). Due to the high capitalcosts for an electron accelerator (usually> 1 million US-$), therelated risk potential from X-rays and hence the necessary securitymeasures, further development of the electron beam process doesnot seem profitable.

3. Oxidation by-products

Oxidation by-product generation during the application of AOPsis a critical factor for process viability. The abundance of nitrogen,halogens and dissolved organic matter (DOM) during disinfectionof AOPs might lead to formation of organic halogenated by-products such as total organic halides (TOX), trihalomethanes(THM), haloacetic acids (HAA), haloacetonitriles (HAN), and inor-ganic by-products (e.g., chlorate, perchlorate and bromate). AllAOPs are based on radical oxidation paths. However, OBP formationis diverse depending on radical type (e.g., OH-, sulfate- or chlorine-radical), radical exposure, abundancy of other influencing waterconstituents (e.g. radical scavengers), and direct reactions ofapplied oxidants such as for instance ozone or chlorine. Occurrenceand health risks of by-products in drinking water as well as anoverview of regulations and guidelines for specific disinfection by-products is comprehensively reviewed (Richardson et al., 2007;Stalter et al., 2016). While assessing the health risks from OBPs ischallenging, there is a general desire to minimize their formation.Thus, in the following sections OBP formation is discussed

considering main reaction mechanisms divided into reactions withinorganic and organic compounds to identify opportunities tominimize their formation.

3.1. Reactions with inorganic compounds

The oxyhalides (chlorite, chlorate, perchlorate and bromate) arepotential inorganic by-products of oxidation processes. Bromate(BrO3

�) which is regulated in drinking water worldwide, cangenerally be formed in a pure �OH reaction if Br� is abundant in thefeed water (von Gunten and Oliveras, 1998). However, this reactionis suppressed by DOM (Lutze et al., 2014) and in processes withexcess H2O2 (UV/H2O2, Fenton reaction), where the oxidation of theintermediate HOBr to BrO3

� is hindered by the fast reductive reac-tion with H2O2 to Br� (k¼ 7.6 $ 108M�1 s�1) (von Gunten, 2003b).Therefore, BrO3

� formation is negligible in most �OH-dominatedsystems. Chlorate and perchlorate formation during OH-radicalprocesses only occurs under certain conditions. The initial reac-tion of OH-radicals with Cl� is slowwith a rate constant in the orderof 103M�1 s�1 at pH 7 and formation of Cl-radicals can therefore beneglected at neutral conditions (von Gunten, 2003b). If oxidativechlorine species are abundant (e.g., ClO�, OCl�), however, sequentialoxidation by OH-radicals to chlorate and perchlorate is possible. InSO4�

- based processes, BrO3� formation may be formed in a direct

reaction of Br� with SO4�- (kSO4�-¼ 3.5 $ 109M�1 s�1 (Redpath and

Willson, 1975)) in the absence of DOM (Lutze et al., 2014). Inaddition, BrO3

� formation may occur from SO4�- reactionwith Cl� to

form �OH and Cl� turning SO4�- based process into a Cl� and �OH

dominated process.In some cases, �OH are directly formed on active surfaces (e.g.

the anode surface in eAOPs or the catalyst surface in heterogeneouscAOPs) and can only react within the diffusion limited zone ofabout <1 mm (Kapałka et al., 2009). These conditions might inducehigh radical densities at the surface allowing kinetically unfavoredoxidation of Cl�, Br� and intermediate species forming bromate,chlorate and perchlorate (e.g. Bergmann and Rollin, 2007). Adetailed literature review on OBP formation in eAOPs has recentlybeen compiled by Chaplin (2014).

In ozone based AOPs at elevated bromide concentrations(>100 mg/L), direct reaction of ozone can lead to 5e50% conversionof bromide to bromate, depending on ozone exposure. The fastdecomposition of ozone in AOPs limits this reaction (von Gunten,2003a), but significant bromate formation was still observed atelevated ozone dosages in the O3/H2O2 process (Hübner et al.,2015). Since chloride is not oxidized by ozone, chlorate formationis only relevant for ozonation, if pre-oxidation by reactive chlorinespecies is applied (von Gunten, 2003b). As discussed above, for-mation of chlorate, perchlorate and bromate is not critical in UVbased AOPs, but other inorganic by-products can be formed. IfDOM-containing water samples (10 mgC/L) are exposed to vacuumUV or LP-UV irradiation, H2O2 formation can reach up to 1.5 and0.3mg/L, respectively (Buchanan et al., 2006). Furthermore,photolysis of nitrate may form nitrite during UV irradiation. Whilethe molar absorption coefficient of nitrate at l¼ 254 nm is low(ε¼ 4M�1cm�1), it increases dramatically below 240 nm (Sharplessand Linden, 2001) revealing higher nitrite formation potentials forMP and VUV systems.

3.2. Reactions with organic compounds

Reactions of �OH with DOM generally involve hydrogenabstraction, electrophilic addition and radical combination. Despitetheir electrophilic character, these reactions are quite diverse andformation of significant OBP concentrations was not observed in�OH dominated oxidation processes for general water applications.


UV/H2O2, for example, was described as an AOP without significantOBP formation and no or minor increase of genotoxic activity ifapplied to surface water (Linden et al., 2005). Some studies, how-ever, describe significant organic OBP formation during UV/H2O2 athigh pH and Cl� concentrations (>1 g/L) evaluated as adsorbableorganic halides (AOX) (Baycan et al., 2007). Despite the slow reac-tivity of �OH with Cl�, at high concentrations of Cl� formation ofchlorine radicals (Cl�, Cl2�) and active chlorine species (e.g., Cl2,OCl�, HOCl) is sufficient. These species can subsequently react withorganic compounds by addition and substitution reactions result-ing in halogenated OBPs (Oppenl€ander, 2003). Main formationpaths of halogenated organic OBPs (e.g., THM, HAN and HAA) arebased on reactions between oxoacids/hypohalites (HOX/OX�) andDOM, where the addition of halogens to DOM increases followingthe order Cl< I< Br (von Gunten, 2003b). Consequently, UV/chlo-rine process may involve formation of AOX. In contrast, under masstransfer limited conditions organic OBPs are continuously gener-ated (Bagastyo et al., 2012).

In ozone based AOPs, the main pathway for generation ofhalogenated organic compounds is still the reaction of HOX/OX�

with DOM as described above. Formation of bromo-organic com-pounds (<10 mg/L depending on bromide concentration) duringraw surface water ozonation has been confirmed by Huang et al.(2005). Since the oxidation of HOI and OI� by ozone is fast(pH< 8), reaction of HOI with DOM can be neglected for ozonationprocesses (von Gunten, 2003b). Formation of HOCl� is not relevantfor ozonation based on the low reactivity of ozone with chloride, asdescribed in section 4.1.1 resulting in a low relevance of chlorinatedorganic OBPs. Of higher concern is the formation of N-Nitro-sodimethylamine (NDMA), a highly carcinogenic substance mainlyformed by chlorination of nitrogen- and organic carbon-containingwaters. Some studies reported its occurrence after ozonation(Andrzejewski et al., 2008). Detailed reaction pathways for NDMAformation during ozonation are proposed by Yang et al. (2009).However, NDMA is not a major by-product of ozonation (vonGunten, 2003b; Andrzejewski et al., 2008). During UV/chlorineNDMA formation could be inhibited in a pilot-scale system byquenching excess chlorine with thiosulfate (Sichel et al., 2011).

In UV-based processes, NDMA is effectively removed by UV-photolysis (Stefan and Bolton, 2002; Mitch et al., 2003; Sharplessand Linden, 2003). Other AOPs are less efficient removing NDMAdue to moderate and low second-order rate constants with �OH(kOH¼ 3.8� 108M�1s�1 (Wols and Hofman-Caris, 2012) and ozone(kO3¼ 0.052M�1s�1 (Lee et al., 2007)). A common strategy tocontrol NDMA concentrations inwater is the removal of precursors,such as dimethylamine, which can easily be oxidized by ozonationor AOP (Lee et al., 2007). However, also subsequent biological stepsare effective to mitigate NDMA (Drewes et al., 2006).

Recently, researchers have shown the potential of mutagenicorganic by-product formation during application of medium-pressure UV irradiation to water containing nitrate (Hofman-Cariset al., 2015; Kolkman et al., 2015). The photolysis products of ni-trate (mainly peroxynitrite) react with DOM by hydroxylation,nitration and nitrosation reactions forming organic OBPs (Martijnet al., 2014). While potential reaction mechanisms have been pro-posed (Reckhow et al., 2010; Shah et al., 2011), a comprehensiveunderstanding has not yet been completed. However, nitrated ar-omatic compounds are expected to be the most toxic OBPs formedin this process (Martijn et al., 2014).

4. Comparison of advanced oxidation processes

EEO values derived for a specific AOP are depending on themolecular structure, the physico-chemical characteristics, such asspecific reaction rate constants, and the concentration range of the

respective contaminant (only if> 1mg/L). Furthermore, watermatrix, process capacity and energy independent process param-eters (e.g., oxidant or catalyst dose) can have a significant influenceon the efficiency of the process. In general, EEO values should onlybe determined for an AOP which is optimized with respect tooxidant demand, reactor geometry, and other process-specific pa-rameters. All these interdependencies should be kept inmindwhilecomparing different AOP technologies via EEO. Hence, the boundaryconditions inwhich EEO values were determined are very importantfor the overall comparison of AOPs via EEO.

It should be noted that additional energy demand for chemicalsor catalysts is not reflected within this figure of merit. The demandfor auxiliary oxidants (e.g. H2O2), however, can be reflected withinthe EEO concept by regarding H2O2 as “stored electric energy”(Rosenfeldt et al., 2006). For example, Müller et al. (2001) calcu-lated an equivalent of 10 kWh for 1 kg of H2O2 (100%) based on thecommercial prices in Germany. However, the majority of publishedEEO values are limited to the electricity which is directly used in theprocess, e.g. for ozone generation or UV-lamp operation.

Several peer-reviewed journal articles deal with the directcomparison of different AOPs in a defined experimental setup withcontrolled conditions, i.e. in terms of water quality to be treated,target contaminant and other process conditions, with the aim toreveal the most efficient AOP technology. However, many of thesestudies are lacking important information, neglecting relevant pa-rameters or testing removal of substances with specific reactivity tooxidants other than �OH, e.g. ozone-reactive or photolyticallydegradable compounds. Furthermore, comparison is only con-ducted in few water matrices and generalization of those resultsand their transfer to application with other water types, contami-nants etc. should therefore be made with careful consideration ofthe respective conditions.

For this reason, we critically reviewed and compared reportedEEO values from different AOPs. Results are discussed and put incontext to studies showing direct comparison, if available. Inaddition, we analyzed major influencing factors on EEO determi-nation based on literature data for the UV/H2O2 process.

4.1. Comparative screening of EEO values for different AOPs

EEO values for numerous AOPs from literature data are illus-trated in Fig. 2 as box plots sorted according to their respectivemedian values. A summary of all data including specific informa-tion on water type, system size and measured compounds is givenat Mendeley Data (https://doi.org/10.17632/n7h8kb4dfh.2). Onlydata meeting the following criteria were included in the figure:

- Incorporated data is published in a peer-reviewed process- Manufacturer data and data from non-peer reviewed sourcesare included if detailed information about the experimentalsetup is given

- If kinetic data is available, compounds, which are susceptible todirect oxidation by e.g. ozone or UV-photolysis will not beregarded. Threshold values for rate constants in O3- and UV-based processes are set at kO3 < 10M�1s�1 and kUV < 10�5m2/J

Data evaluation was conducted in three steps: screening of EEOvalues, single outlier detection and removal and descriptive sta-tistics. Outlier detectionwas performed by the Dixon test assuminglog-normal distribution for all data sets using an online tool avail-able at http://contchart.com/outliers.aspx. Significance testing wasperformed using the two-sample t-test provided by another onlinetool (http://www.evanmiller.org/ab-testing/t-test.html) assuminglog-normal distribution for all data-sets.

Reported EEO values for individual AOPs often vary by several

https://doi.org/10.17632/n7h8kb4dfh.2

http://contchart.com/outliers.aspx

http://www.evanmiller.org/ab-testing/t-test.html

Fig. 2. Overview of published EEO-values of different AOPs sorted according to median values. For O3- and UV-based AOP data, only substances resistant to direct ozonation/photolysis are shown (references are shown in Table S1). Median values and number of data points are reported on the second and third y-axis, respectively.

Fig. 3. Overview of published EEO-values of the UV/H2O2 process, classified into lab-,pilot- and full-scale data. n refers to the number of data points behind the boxplots.Median values are reported next to the 50th percentile line.


orders of magnitude. In case of ozonation, the strong variabilitymight be explained by the dependence of radical formation fromwater matrix, since it is only initiated by hydroxide ions at elevatedpH or from ozone reactions with organic matter. Little variability ofother processes might either indicate lower sensitivity to waterquality and system design or limited experimental differences inliterature data, e.g. oxidation with microwaves was only tested inultrapure water.

Despite cases of high variability, significant differences betweenAOPs can be observed from the literature study. Based on medianvalues, AOPs are classified in three groups: processes with medianEEO values< 1 kWh/m3 (O3, O3/H2O2, O3/UV, UV/H2O2, UV/persul-fate, UV/chlorine and electron beam) represent a realistic range forfull-scale application (group 1). Photo-Fenton, plasma and eAOPwith median EEO values of 2.6, 3.3, and 38.1 kWh/m3 (1e100 kWh/m3), respectively, are likely too energy intensive for most practicalapplications (group 2). However, they might still provide attractivesolutions for specific challenges and full-scale applicability of theseprocesses should be further investigated. EEO values for group 2processes are significantly higher than values of group 1(p¼ 0.045). Processes with E�EO values> 100 kWh/m3, i.e. UV-basedphotocatalysis, ultrasound and microwave-based AOPs represent-ing high median values of 335, 2616 and 543 kWh/m3, respectively,are considered as not (yet) energy efficient AOPs. Significance ofdifference between group 2 and 3 is calculated as p¼ 0.002.

There are only few quality studies directly comparing processesfrom these different groups. A direct comparison of UV/H2O2 andBDD treatment for aniline removal from synthetic wastewater so-lutions confirmed lower EEO values for UV/H2O2 by about 30%(Benito et al., 2017). However, extensive concentrations of H2O2were applied in this study (1e5 g/L) that might have influencedcalculated EEO values of UV/H2O2 by re-scavenging of �OH by H2O2.

Observed differences between AOPs in the first group are sta-tistically not significant (p>0.2 in between all AOPs) andmost likelydepend on experimental conditions. This is confirmed by severalstudies comparing O3/H2O2 and UV/H2O2. Sutherland and co-workers published a comprehensive evaluation of MTBE oxidation

from five contaminated groundwaters with highly variable waterquality characteristics (Sutherland et al., 2004). Depending onwater type and adjusted pH, either O3/H2O2 or UV/H2O2 achievedlower EEO values. In contrast, Lester et al. (2011) reported lowest E�EOvalues for O3/H2O2 followed by O3, UV/H2O2/O3, UV/O3 and UV forpharmaceutical degradation in phosphate buffer. Also, Müller et al.(2001) showed advantages of the O3/H2O2 combination in com-parison to UV/H2O2 and UV/O3. A recent study compared the effi-ciency of UV/PDS and UV/H2O2 during iodoacids degradation,considering most relevant influencing factors (e.g. photo-susceptibility and process capacity) (Xiao et al., 2016). Specialemphasis was directed to water matrix and oxidant dose effect.


Results revealed higher energy efficiencies for the sulfate radicalbased AOP by a factor of >3. However, comparison of these pro-cesses using single compounds or specific compound groups needsto be evaluated carefully since sulfate radicals react more selec-tively than �OH.

4.2. Principal influences on EEO-values shown at the UV/H2O2

process

Most boxplots for individual AOPs in Fig. 2 reveal high variances,being mainly based on data variability considered in this review.Water quality, process capacity and the selection of target chem-icals are key parameters, which may lead to large deviations of EEOvalues within one process. Effects of these factors is exemplarilyillustrated in the following section using the dataset of UV/H2O2,which is one of the most intensively investigated AOPs and there-fore supplies high data density of EEO values (n¼ 149, excludingsingle outliers and data of photo-susceptible compounds). How-ever, transferability of influencing factors is not ensured for otherAOPs since most influences are process specific.

4.2.1. Influence of process capacityEEO values from UV/H2O2 studies conducted at laboratory, pilot

and full scale are illustrated in Fig. 3. Median EEO values decreasewith process capacity from 2.2 kWh/m3 at lab-scale to 0.68 and0.5 kWh/m3 for pilot- and full-scale applications, respectively. Asignificant difference could be observed (p< 0.05) between lab-and pilot-scale data. The median values indicate that up-scalingenhances energy efficiency, which confirms the findings fromBolton and Stefan (2002). Results furthermore emphasize thatcomparison of lab-scale energy consumption not necessarily rep-resents operation at full-scale. If possible, energy demand shouldrather be estimated based on full-scale system designwith relevantoperational parameters (i.e., oxidant dosage, UV-fluence) deter-mined in standardized lab- or pilot-scale experiments.

4.2.2. Influence of compound reactivityEnergy efficiency of AOPs is also dependent on compound

reactivity. Therefore, gathered EEO values from reviewed UV/H2O2publications were correlated to second-order rate constants (kOH)of photo-resistant target chemicals. Results indicated a slightly

Fig. 4. Reviewed EEO values of the UV/H2O2 process: Effect of different water matrixapplications; n refers to the number of data points behind the boxplots. Median valuesare reported next to the 50th percentile line.

negative correlation (R2¼ 0.21), confirming, that substances withhigher kOH values are more efficiently oxidized with a lower energyeffort (Fig. S1). However, due to the low correlation coefficient theinfluence of other parameters (e.g. process capacity) is assumed tobe higher.

4.2.3. Influence of water qualityWater quality mainly affects the UV/H2O2 process by UV-

transmittance and radical scavengers. Therefore, EEO values wereinvestigated based on water characteristics reported from therespective article. Since numerous water types were included in alldata gathered, ranging from ultrapure lab water to industrialwastewater effluents, EEO values were classified into main waterapplication groups: pure water, drinking water, groundwater andwastewater applications. Pure water applications include lab-scaleexperiments with deionized and ultrapure water but also pilot-scale applications with reverse osmosis permeate (i.e. in waterreuse). Drinking water applications summarize UV/H2O2 processeswith surface water after pre-treatment with various process com-binations. Groundwater applications include AOPs at contaminatedsites as well as drinking water applications from groundwater.Wastewater consists of secondary and tertiary effluent frommunicipal wastewater treatment plants and industrial applications.Resulting EEO values of pure water, drinking water, groundwater,and wastewater applications are presented in Fig. 4 as box plots.Median EEO values of each applicationwere determined as 2.7, 0.63,2.7 and 2.2 kWh/m3, respectively. However, no significant differ-ence could be observed between different groups (p> 0.05). Theconcentration of radical scavengers and UV absorbance (UVA) arethe most influential parameters for radical yield and radicaloxidation efficiency. Consequently, waters with higher scavengerconcentrations (and different scavenger composition) and higherUVA should result in higher EEO values. Surprisingly, ultrapurewater applications reveal the highest median EEO value in thiscomparison. A possible explanation for this finding might be thepredominant use of pure water in lab-scale experiments whichbiases the illustration as already discussed in section 4.2.1. Datacomparison from drinking and wastewater applications revealhigher energy needs with increasing scavenger content. Especially,DOC and consequently UVA are expected to be higher in this order.

Fig. 5. Overview of published EEO-values of the UV/H2O2 process, classified into databased on medium- and low-pressure lamps. Headline shows the number of data pointsbehind the boxplots. Median values are reported next to the 50th percentile line.


In contrast to drinking water, the median EEO value from ground-water applications is similar towastewater oxidation. Groundwatermay contain strongly variable inorganic concentrations, e.g. alka-linity (HCO3

�/CO32�), but also reduced species like manganese and

iron, which might significantly scavenge OH-radicals.Overall, the operational classification of AOPs presented in Fig. 4

did not suggest any significant effects of water matrix on EEO values,probably because the selected categories like pure water, drinkingwater, groundwater and wastewater are not specific enough andcan include a wide range of different applications and water qual-ities. Therefore, a direct correlation of water quality parameterswith reported EEO values was also investigated. Considered as mostrelevant parameters, DOC, UV transmittance and turbidity (re-ported as NTU) data were provided for 124, 131 and 31 of thereviewed 147 data sets, respectively. A direct relationship withinthe reviewed data set between EEO values and DOC concentrations,UV transmittance or NTU, however, could not be revealed (Figs. S2and S3).

4.2.4. Influence of lamp typeThe influence of different lamp types on EEO values is illustrated

in Fig. 5. MP UV lamps result in significantly higher EEO valuescompared to LP lamps (p< 0.001). The respective median valuescan be determined as 1.0 and 0.4 kWh/m3. This is not surprising,since the molar absorption coefficient of H2O2 increases at wave-lengths <260 nm and LP lamps depict higher (<35% at 254 nm)energy efficiencies than MP lamps (<10% at 254 nm). Inevitably, LPlamps yield in a higher H2O2 activation and consequently in lowerEEO values. This was also confirmed by Rosenfeldt et al. (2005), whodirectly compared LP and MP UV lamps for the oxidation of 2-methyl-isoborneol (2-MIB) and geosmin. Raw blend surface waterand filtered clearwell water were used in lab-scale reactors. EEOvalues revealed that LP lamps can be more energy efficient than MPlamps for �OH generation.

5. Conclusions

This study provides a critical review of different established andemerging AOPs including a mechanistic discussion of process-specific by-product formation. To facilitate a comparison of en-ergy efficiency, datawere collected for various AOPs in an extensiveanalysis of peer-reviewed journal articles and critically comparedbased on reported EEO values. Despite high variability of resultsfrom individual processes, significant differences between AOPsefficiency were observed. Based on reported EEO values, processeswere classified into (1) AOPs with median EEO values of <1 kWh/m3

(O3, O3/H2O2, O3/UV, UV/H2O2, UV/persulfate, UV/chlorine, electronbeam), (2) processes with median EEO values in the range of 1e100kWh/m3 (Photo-Fenton, plasma, and electrolytic AOPs) and (3) UV-based photocatalysis, ultrasound, and microwave-based AOPs(median EEO values of >100 kWh/m3), which are considered as not(yet) energy efficient AOPs. A more detailed evaluation of data forthe UV/H2O2 process showed highest impact of UV-lamp type,water matrix, and process capacity (lab-scale vs. pilot- and full-scale) on resulting EEO values. No significant correlation could beobserved between EEO values and compound reactivity with OH-radicals. In addition, reviewed literature indicates that by-productformation from hydroxyl radicals is not critical unless formed athigh density on surface areas (e.g. in electrolytic AOPs). However,AOPs involving other oxidants such as ozone, sulfate radicals orchlorine radicals need to be evaluated in more detail since site- andprocess-specific by-products might be formed.

This study confirmed the main limitation to use the EEO conceptfor a general comparison of different AOPs due to the variability ofthe above-mentioned influencing factors. However, if all factors are

considered within a direct comparison, the EEO concept provides apowerful figure of merit to directly compare and evaluate AOPsbased on energy efficiency.

Appendix A. Supplementary data

Supplementary data related to this article can be found athttps://doi.org/10.1016/j.watres.2018.03.042.

References

Ahn, Y., Lee, D., Kwon, M., Choi, I.-H., Nam, S.-N., Kang, J.-W., 2017. Characteristicsand fate of natural organic matter during UV oxidation processes. Chemosphere184, 960e968. https://doi.org/10.1016/j.chemosphere.2017.06.079.

Alaton, I.A., Balcioglu, I.A., Bahnemann, D.W., 2002. Advanced oxidation of a reactivedyebath effluent: comparison of O3, H2O2/UV-C and TiO2/UV-A processes.Water Res. 36 (5), 1143e1154. https://doi.org/10.1016/S0043-1354(01)00335-9.

Andrzejewski, P., Kasprzyk-Hordern, B., Nawrocki, J., 2008. N-nitrosodimethylamine(NDMA) formation during ozonation of dimethylamine-containing waters.Water Res. 42 (4e5), 863e870. https://doi.org/10.1016/j.watres.2007.08.032.

Antoniou, M.G., La Cruz, A.A. de, Dionysiou, D.D., 2010. Degradation of microcystin-LR using sulfate radicals generated through photolysis, thermolysis and e�transfer mechanisms. Appl. Catal. B Environ. 96 (3e4), 290e298. https://doi.org/10.1016/j.apcatb.2010.02.013.

Ao, X., Liu, W., 2016. Degradation of sulfamethoxazole by medium pressure UV andoxidants: peroxymonosulfate, persulfate, and hydrogen peroxide. Chem. Eng. J.https://doi.org/10.1016/j.cej.2016.12.089.

Audenaert, W.T.M., Vermeersch, Y., van Hulle, S.W.H., Dejans, P., Dumoulin, A.,Nopens, I., 2011. Application of a mechanistic UV/hydrogen peroxide model atfull-scale: sensitivity analysis, calibration and performance evaluation. Chem.Eng. J. 171 (1), 113e126. https://doi.org/10.1016/j.cej.2011.03.071.

Autin, O., Romelot, C., Rust, L., Hart, J., Jarvis, P., MacAdam, J., Parsons, S.A.,Jefferson, B., 2013. Evaluation of a UV-light emitting diodes unit for the removalof micropollutants in water for low energy advanced oxidation processes.Chemosphere 92 (6), 745e751. https://doi.org/10.1016/j.chemosphere.2013.04.028.

Bae, W., Won, H., Hwang, B., Toledo, R.A. de, Chung, J., Kwon, K., Shim, H., 2015.Characterization of refractory matters in dyeing wastewater during a full-scaleFenton process following pure-oxygen activated sludge treatment. J. HazardMater. 287, 421e428. https://doi.org/10.1016/j.jhazmat.2015.01.052.

Bagastyo, A.Y., Batstone, D.J., Kristiana, I., Gernjak, W., Joll, C., Radjenovic, J., 2012.Electrochemical oxidation of reverse osmosis concentrate on boron-dopeddiamond anodes at circumneutral and acidic pH. Water Res. 46 (18),6104e6112. https://doi.org/10.1016/j.watres.2012.08.038.

Bai, Z., Yang, Q., Wang, J., 2016. Catalytic ozonation of sulfamethazine antibioticsusing Ce0.1Fe0.9OOH: catalyst preparation and performance. Chemosphere 161,174e180. https://doi.org/10.1016/j.chemosphere.2016.07.012.

Baycan, N., Thomanetz, E., Sengül, F., 2007. Influence of chloride concentration onthe formation of AOX in UV oxidative system. J. Hazard Mater. 143 (1e2),171e176. https://doi.org/10.1016/j.jhazmat.2006.09.010.

Bejan, D., Malcolm, J.D., Morrison, L., Bunce, N.J., 2009. Mechanistic investigation ofthe conductive ceramic Ebonex® as an anode material. Electrochim. Acta 54(23), 5548e5556. https://doi.org/10.1016/j.electacta.2009.04.057.

Beltr�an, F.J., 2004. Ozone Reaction Kinetics for Water and Wastewater Systems.Lewis Publishers, Boca Raton, Fla.

Benito, A., Penad�es, A., Lliberia, J.L., Gonzalez-Olmos, R., 2017. Degradation pathwaysof aniline in aqueous solutions during electro-oxidation with BDD electrodesand UV/H2O2 treatment. Chemosphere 166, 230e237. https://doi.org/10.1016/j.chemosphere.2016.09.105.

Bergmann, M.E.H., Iourtchouk, T., Rollin, J., 2011. The occurrence of bromate andperbromate on BDD anodes during electrolysis of aqueous systems containingbromide: first systematic experimental studies. J. Appl. Electrochem. 41 (9),1109e1123. https://doi.org/10.1007/s10800-011-0329-5.

Bergmann, M.H., Rollin, J., 2007. Product and by-product formation in laboratorystudies on disinfection electrolysis of water using boron-doped diamond an-odes. Catal. Today 124 (3e4), 198e203. https://doi.org/10.1016/j.cattod.2007.03.038.

Blake, D.M., 2001. Bibliography of Work on the Heterogeneous PhotocatalyticRemoval of Hazardous Compounds from Water and Air (NREL/TP-510-31319).National Renewable Energy Laboratory, Colorado.

Bo, L., Quan, X., Chen, S., Zhao, H., Zhao, Y., 2006. Degradation of p-nitrophenol inaqueous solution by microwave assisted oxidation process through a granularactivated carbon fixed bed. Water Res. 40 (16), 3061e3068. https://doi.org/10.1016/j.watres.2006.06.030.

Bokare, A.D., Choi, W., 2014. Review of iron-free Fenton-like systems for activatingH2O2 in advanced oxidation processes. J. Hazard Mater. 275, 121e135. https://doi.org/10.1016/j.jhazmat.2014.04.054.

Bolton, J.R., Bircher, K.G., Tumas, W., Tolman, C.A., 1996. Figures-of merit for thetechnical development and application of advanced oxidation processes. J. Adv.Oxid. Technol. 1, 13e17. https://doi.org/10.1515/jaots-1996-0104.

Bolton, J.R., Bircher, K.G., Tumas, W., Tolman, C.A., 2001. Figures-of-merit for the


https://doi.org/10.1016/j.chemosphere.2017.06.079

https://doi.org/10.1016/S0043-1354(01)00335-9


https://doi.org/10.1016/j.apcatb.2010.02.013


https://doi.org/10.1016/j.cej.2016.12.089




https://doi.org/10.1016/j.jhazmat.2015.01.052




https://doi.org/10.1016/j.electacta.2009.04.057

http://refhub.elsevier.com/S0043-1354(18)30238-0/sref141





https://doi.org/10.1007/s10800-011-0329-5

https://doi.org/10.1016/j.cattod.2007.03.038

https://doi.org/10.1016/j.cattod.2007.03.038








https://doi.org/10.1515/jaots-1996-0104


technical development and application of advanced oxidation technologies forboth electric- and solar-driven systems (IUPAC Technical Report). Pure Appl.Chem. 73 (4). https://doi.org/10.1351/pac200173040627.

Bolton, J.R., Stefan, M.I., 2002. Fundamental photochemical approach to the con-cepts of fluence (UV dose) and electrical energy efficiency in photochemicaldegradation reactions. Res. Chem. Intermed. 28 (7e9), 857e870. https://doi.org/10.1163/15685670260469474.

Bolton, J.R., Valladares, J.E., Zanin, J.P., Cooper, W.J., Nickelsen, M.G., Kajdi, D.C.,Waite, Kurucz, C.N., 1998. Figures-of-Merit for advanced oxidation technolo-gies: a comparison of homogeneous UV/H2O2, heterogeneous UV/TiO2 andelectron beam processes. J. Adv. Oxid. Technol. 3 (2), 174e181. https://doi.org/10.1515/jaots-1998-0211.

Bonfatti, F., 1999. Electrochemical incineration of glucose as a model organic sub-strate. I. Role of the electrode material. J. Electrochem. Soc. 146 (6), 2175.https://doi.org/10.1149/1.1391909.

Buchanan, W., Roddick, F., Porter, N., 2006. Formation of hazardous by-productsresulting from the irradiation of natural organic matter: comparison betweenUV and VUV irradiation. Chemosphere 63 (7), 1130e1141. https://doi.org/10.1016/j.chemosphere.2005.09.040.

Buffle, M.-O., Schumacher, J., Meylan, S., Jekel, M., von Gunten, U., 2006. Ozonationand advanced oxidation of wastewater: effect of O3 dose, pH, DOM and HO�

-scavengers on ozone decomposition and HO� generation. Ozone: Sci. Eng. 28(4), 247e259. https://doi.org/10.1080/01919510600718825.

Buffle, M.-O., von Gunten, U., 2006. Phenols and amine induced HO� generationduring the initial phase of natural water ozonation. Environ. Sci. Technol. 40 (9),3057e3063. https://doi.org/10.1021/es052020c.

Buxton, G.V., Greenstock, C.L., Helman, W.P., Ross, A.B., Tsang, W., 1988. Criticalreview of rate constants for reactions of hydrated electrons, hydrogen atomsand hydroxyl radicals (�OH/�O�) in aqueous solution. J. Phys. Chem. Ref. Data 17(2), 513. https://doi.org/10.1063/1.555805.

Calderara, V., Jekel, M., Zaror, C., 2002. Ozonation of 1-naphthalene, 1,5-naphthalene, and 3-nitrobenzene sulphonic acids in aqueous solutions. Envi-ron. Technol. 23 (4), 373e380. https://doi.org/10.1080/09593332508618403.

Chaplin, B.P., 2014. Critical review of electrochemical advanced oxidation processesfor water treatment applications. Environ. Sci. Process. Impacts 16 (6),1182e1203. https://doi.org/10.1039/c3em00679d.

Chaplin, B.P., Hubler, D.K., Farrell, J., 2013. Understanding anodic wear at borondoped diamond film electrodes. Electrochim. Acta 89, 122e131. https://doi.org/10.1016/j.electacta.2012.10.166.

Chen, J., Loeb, S., Kim, J.-H., 2017. LED revolution: fundamentals and prospects forUV disinfection applications. Environ. Sci. Water Res. Technol. 3 (2), 188e202.https://doi.org/10.1039/C6EW00241B.

Collins, J., Bolton, J.R. (Eds.), 2016. Advanced Oxidation Handbook, first ed. AmericanWater Works Association, Denver, CO.

Comninellis, C., Kapalka, A., Malato, S., Parsons, S.A., Poulios, I., Mantzavinos, D.,2008. Advanced oxidation processes for water treatment: advances and trendsfor R&D. J. Chem. Technol. Biotechnol. 83 (6), 769e776. https://doi.org/10.1002/jctb.1873.

Dong, S., Feng, J., Fan, M., Pi, Y., Hu, L., Han, X., Liu, M., Sun, J., Sun, J., 2015. Recentdevelopments in heterogeneous photocatalytic water treatment using visiblelight-responsive photocatalysts: a review. RSC Adv. 5 (19), 14610e14630.https://doi.org/10.1039/C4RA13734E.

Drewes, J.E., Hoppe, C., Jennings, T., 2006. Fate and transport of N-Nitrosaminesunder conditions simulating full-scale groundwater recharge operations. WaterEnviron. Res. 78 (13), 2466e2473. https://doi.org/10.2175/106143006X115408.

Drewes, J.E., Khan, S.J., 2015. Contemporary design, operation, and monitoring ofpotable reuse systems. J. Water Reuse and Desalination 5 (1), 1. https://doi.org/10.2166/wrd.2014.148.

Elovitz, M.S., von Gunten, U., 1999. Hydroxyl radical/ozone ratios during ozonationprocesses. I. The R ct concept. Ozone: Sci. Eng. 21 (3), 239e260. https://doi.org/10.1080/01919519908547239.

EPA, 2006. National Primary Drinking Water Regulations: Long Term 2 EnhancedSurface Water Treatment Rule; Final Rule. Environmental Protection Agency.Federal Register 40 CFR Parts 9, 141, and 142.

Fang, J., Fu, Y., Shang, C., 2014. The roles of reactive species in micropollutantdegradation in the UV/free chlorine system. Environ. Sci. Technol. 48 (3),1859e1868. https://doi.org/10.1021/es4036094.

Fast, S.A., Gude, V.G., Truax, D.D., Martin, J., Magbanua, B.S., 2017. A critical evalu-ation of advanced oxidation processes for emerging contaminants removal.Environ. Process. 4 (1), 283e302. https://doi.org/10.1007/s40710-017-0207-1.

Fernandes, A., Santos, D., Pacheco, M.J., Ciríaco, L., Lopes, A., 2014. Nitrogen andorganic load removal from sanitary landfill leachates by anodic oxidation at Ti/Pt/PbO2, Ti/Pt/SnO2-Sb2O4 and Si/BDD. Appl. Catal. B Environ. 148-149,288e294. https://doi.org/10.1016/j.apcatb.2013.10.060.

Ghazi, N.M., Lastra, A.A., Watts, M.J., 2014. Hydroxyl radical (OH) scavenging inyoung and mature landfill leachates. Water Res. 56, 148e155. https://doi.org/10.1016/j.watres.2014.03.001.

Giannakis, S., Gamarra Vives, F.A., Grandjean, D., Magnet, A., Alencastro, L.F. de,Pulgarin, C., 2015. Effect of advanced oxidation processes on the micro-pollutants and the effluent organic matter contained in municipal wastewaterpreviously treated by three different secondary methods. Water Res. 84,295e306. https://doi.org/10.1016/j.watres.2015.07.030.

Goel, M., Hongqiang, H., Mujumdar, A.S., Ray, M.B., 2004. Sonochemical decom-position of volatile and non-volatile organic compoundsea comparative study.Water Res. 38 (19), 4247e4261. https://doi.org/10.1016/j.watres.2004.08.008.

Gros, M., Petrovi�c, M., Ginebreda, A., Barcel�o, D., 2010. Removal of pharmaceuticalsduring wastewater treatment and environmental risk assessment using hazardindexes. Environ. Int. 36 (1), 15e26. https://doi.org/10.1016/j.envint.2009.09.002.

Guan, Y.-H., Ma, J., Li, X.-C., Fang, J.-Y., Chen, L.-W., 2011. Influence of pH on theformation of sulfate and hydroxyl radicals in the UV/peroxymonosulfate sys-tem. Environ. Sci. Technol. 45 (21), 9308e9314. https://doi.org/10.1021/es2017363.

Haenni, W., Baumann, H., Comninellis, C., Gandini, D., Niedermann, P., Perret, A.,Skinner, N., 1998. Diamond-sensing microdevices for environmental controland analytical applications. Diam. Relat. Mater. 7 (2e5), 569e574. https://doi.org/10.1016/S0925-9635(97)00253-7.

Han, D.-H., Cha, S.-Y., Yang, H.-Y., 2004. Improvement of oxidative decomposition ofaqueous phenol by microwave irradiation in UV/H2O2 process and kineticstudy. Water Res. 38 (11), 2782e2790. https://doi.org/10.1016/j.watres.2004.03.025.

Hijosa-Valsero, M., Molina, R., Bayona, J.M., 2014. Assessment of a dielectric barrierdischarge plasma reactor at atmospheric pressure for the removal of bisphenolA and tributyltin. Environ. Technol. 35 (9e12), 1418e1426. https://doi.org/10.1080/09593330.2013.869624.

Hislop, K.A., Bolton, J.R., 1999. The photochemical generation of hydroxyl radicals inthe UV-vis/Ferrioxalate/H2O2 system. Environ. Sci. Technol. 33 (18), 3119e3126.https://doi.org/10.1021/es9810134.

Hofman-Caris, R.C.H.M., Harmsen, D.J.H., Puijker, L., Baken, K.A., Wols, B.A.,Beerendonk, E.F., Keltjens, L.L.M., 2015. Influence of process conditions andwater quality on the formation of mutagenic byproducts in UV/H2O2 processes.Water Res. 74, 191e202. https://doi.org/10.1016/j.watres.2015.01.035.

Huang, W.-J., Fang, G.-C., Wang, C.-C., 2005. The determination and fate of disin-fection by-products from ozonation of polluted raw water. Sci. Total Environ.345 (1e3), 261e272. https://doi.org/10.1016/j.scitotenv.2004.10.019.

Hübner, U., Zucker, I., Jekel, M., 2015. Options and limitations of hydrogen peroxideaddition to enhance radical formation during ozonation of secondary effluents.J. Water Reuse and Desalination 5 (1), 8. https://doi.org/10.2166/wrd.2014.036.

Huerta-Fontela, M., Galceran, M.T., Ventura, F., 2010. Fast liquid chromatography-quadrupole-linear ion trap mass spectrometry for the analysis of pharmaceu-ticals and hormones in water resources. J. Chromatogr., A 1217 (25), 4212e4222.https://doi.org/10.1016/j.chroma.2009.11.007.

Ike, I.A., Linden, K.G., Orbell, J.D., Duke, M., 2018. Critical review of the science andsustainability of persulphate advanced oxidation processes. Chem. Eng. J.https://doi.org/10.1016/j.cej.2018.01.034.

Jiang, B., Zheng, J., Qiu, S., Wu, M., Zhang, Q., Yan, Z., Xue, Q., 2014. Review onelectrical discharge plasma technology for wastewater remediation. Chem. Eng.J. 236, 348e368. https://doi.org/10.1016/j.cej.2013.09.090.

Jiang, C., Pang, S., Ouyang, F., Ma, J., Jiang, J., 2010. A new insight into Fenton andFenton-like processes for water treatment. J. Hazard Mater. 174 (1e3), 813e817.https://doi.org/10.1016/j.jhazmat.2009.09.125.

Jin, J., El-Din, M.G., Bolton, J.R., 2011. Assessment of the UV/chlorine process as anadvanced oxidation process. Water Res. 45 (4), 1890e1896. https://doi.org/10.1016/j.watres.2010.12.008.

Kapałka, A., F�oti, G., Comninellis, C., 2009. The importance of electrode material inenvironmental electrochemistry. Electrochim. Acta 54 (7), 2018e2023. https://doi.org/10.1016/j.electacta.2008.06.045.

Kaptijn, J.P., 1997. The ecoclear® process. Results from full-scale installations. OzoneSci. Eng. 19 (4), 297e305. https://doi.org/10.1080/01919519708547294.

Katsoyiannis, I.A., Canonica, S., von Gunten, U., 2011. Efficiency and energy re-quirements for the transformation of organic micropollutants by ozone, O3/H2O2 and UV/H2O2. Water Res. 45 (13), 3811e3822. https://doi.org/10.1016/j.watres.2011.04.038.

Keen, O.S., Love, N.G., Aga, D.S., Linden, K.G., 2016. Biodegradability of iopromideproducts after UV/H2O2 advanced oxidation. Chemosphere 144, 989e994.https://doi.org/10.1016/j.chemosphere.2015.09.072.

Keen, O.S., Love, N.G., Linden, K.G., 2012. The role of effluent nitrate in trace organicchemical oxidation during UV disinfection. Water Res. 46 (16), 5224e5234.https://doi.org/10.1016/j.watres.2012.06.052.

Kesselman, J.M., Weres, O., Lewis, N.S., Hoffmann, M.R., 1997. Electrochemicalproduction of hydroxyl radical at polycrystalline Nb-Doped TiO 2 electrodesand estimation of the partitioning between hydroxyl radical and direct holeoxidation pathways. J. Phys. Chem. B 101 (14), 2637e2643. https://doi.org/10.1021/jp962669r.

Khan, J.A., He, X., Shah, N.S., Khan, H.M., Hapeshi, E., Fatta-Kassinos, D.,Dionysiou, D.D., 2014. Kinetic and mechanism investigation on the photo-chemical degradation of atrazine with activated H2O2, S2O82� and HSO5�.Chem. Eng. J. 252, 393e403. https://doi.org/10.1016/j.cej.2014.04.104.

Klavarioti, M., Mantzavinos, D., Kassinos, D., 2009. Removal of residual pharma-ceuticals from aqueous systems by advanced oxidation processes. Environ. Int.35 (2), 402e417. https://doi.org/10.1016/j.envint.2008.07.009.

Kolkman, A., Martijn, B.J., Vughs, D., Baken, K.A., van Wezel, A.P., 2015. Tracingnitrogenous disinfection byproducts after medium pressure UV water treat-ment by stable isotope labeling and high resolution mass spectrometry. Envi-ron. Sci. Technol. 49 (7), 4458e4465. https://doi.org/10.1021/es506063h.

Kruithof, J.C., Kamp, P.C., Martijn, B.J., 2007. UV/H2O2 treatment: a practical solutionfor organic contaminant control and primary disinfection. Ozone Sci. Eng. 29(4), 273e280. https://doi.org/10.1080/01919510701459311.

Lee, C., Yoon, J., von Gunten, U., 2007. Oxidative degradation of N-nitro-sodimethylamine by conventional ozonation and the advanced oxidation

https://doi.org/10.1351/pac200173040627

https://doi.org/10.1163/15685670260469474

https://doi.org/10.1163/15685670260469474



https://doi.org/10.1149/1.1391909



https://doi.org/10.1080/01919510600718825

https://doi.org/10.1021/es052020c

https://doi.org/10.1063/1.555805

https://doi.org/10.1080/09593332508618403

https://doi.org/10.1039/c3em00679d



https://doi.org/10.1039/C6EW00241B



https://doi.org/10.1002/jctb.1873

https://doi.org/10.1002/jctb.1873

https://doi.org/10.1039/C4RA13734E

https://doi.org/10.2175/106143006X115408

https://doi.org/10.2166/wrd.2014.148

https://doi.org/10.2166/wrd.2014.148

https://doi.org/10.1080/01919519908547239

https://doi.org/10.1080/01919519908547239




https://doi.org/10.1021/es4036094

https://doi.org/10.1007/s40710-017-0207-1






https://doi.org/10.1016/j.envint.2009.09.002


https://doi.org/10.1021/es2017363

https://doi.org/10.1021/es2017363

https://doi.org/10.1016/S0925-9635(97)00253-7

https://doi.org/10.1016/S0925-9635(97)00253-7



https://doi.org/10.1080/09593330.2013.869624

https://doi.org/10.1080/09593330.2013.869624

https://doi.org/10.1021/es9810134


https://doi.org/10.1016/j.scitotenv.2004.10.019

https://doi.org/10.2166/wrd.2014.036

https://doi.org/10.1016/j.chroma.2009.11.007








https://doi.org/10.1080/01919519708547294





https://doi.org/10.1021/jp962669r

https://doi.org/10.1021/jp962669r



https://doi.org/10.1021/es506063h

https://doi.org/10.1080/01919510701459311


process ozone/hydrogen peroxide. Water Res. 41 (3), 581e590. https://doi.org/10.1016/j.watres.2006.10.033.

Legrini, O., Oliveros, E., Braun, A.M., 1993. Photochemical processes for watertreatment. Chem. Rev. 93 (2), 671e698. https://doi.org/10.1021/cr00018a003.

Lester, Y., Avisar, D., Gozlan, I., Mamane, H., 2011. Removal of pharmaceuticals usingcombination of UV/H2O2/O3 advanced oxidation process. Water Sci. Technol.: J.Int. Assoc. Water Pollut. Res. 64 (11), 2230e2238. https://doi.org/10.2166/wst.2011.079.

Lim, M.H., 2008. Fate of Wastewater-derived Contaminants in Surface Waters.Dissertation. University of California, Berkeley.

Linden, K.G., Sharpless, C.M., Andrews, S., Atasi, K., Korategere, V., Stefan, M.,Suffet, I.M., 2005. Innovative UV Technologies to Oxidize Organic and Organ-oleptic Chemicals. Water Environment Research Foundation.

Locke, B.R., Sato, M., Sunka, P., Hoffmann, M.R., Chang, J.-S., 2006. Electrohydraulicdischarge and nonthermal plasma for water treatment. Ind. Eng. Chem. Res. 45(3), 882e905. https://doi.org/10.1021/ie050981u.

Luo, Y., Guo, W., Ngo, H.H., Nghiem, L.D., Hai, F.I., Zhang, J., Liang, S., Wang, X.C.,2014. A review on the occurrence of micropollutants in the aquatic environ-ment and their fate and removal during wastewater treatment. Sci. Total En-viron. 473-474, 619e641. https://doi.org/10.1016/j.scitotenv.2013.12.065.

Lutterbeck, C.A., Machado, E.L., Kümmerer, K., 2015. Photodegradation of the anti-neoplastic cyclophosphamide: a comparative study of the efficiencies of UV/H2O2, UV/Fe2þ/H2O2 and UV/TiO2 processes. Chemosphere 120, 538e546.https://doi.org/10.1016/j.chemosphere.2014.08.076.

Lutze, H., 2013. Sulfate Radical Based Oxidation in Water Treatment. Dissertation.University of Duisburg-Essen, Germany.

Lutze, H.V., Bakkour, R., Kerlin, N., von Sonntag, C., Schmidt, T.C., 2014. Formation ofbromate in sulfate radical based oxidation: mechanistic aspects and suppres-sion by dissolved organic matter. Water Res. 53, 370e377. https://doi.org/10.1016/j.watres.2014.01.001.

Lutze, H.V., Kerlin, N., Schmidt, T.C., 2015. Sulfate radical-based water treatment inpresence of chloride: formation of chlorate, inter-conversion of sulfate radicalsinto hydroxyl radicals and influence of bicarbonate. Water Res. 72, 349e360.https://doi.org/10.1016/j.watres.2014.10.006.

Mahamuni, N.N., Adewuyi, Y.G., 2010. Advanced oxidation processes (AOPs)involving ultrasound for waste water treatment: a review with emphasis oncost estimation. Ultrason. Sonochem. 17 (6), 990e1003. https://doi.org/10.1016/j.ultsonch.2009.09.005.

Mahdi-Ahmed, M., Chiron, S., 2014. Ciprofloxacin oxidation by UV-C activatedperoxymonosulfate in wastewater. J. Hazard Mater. 265, 41e46. https://doi.org/10.1016/j.jhazmat.2013.11.034.

Martijn, A.J., Boersma, M.G., Vervoort, J.M., Rietjens, I.M.C.M., Kruithof, J.C., 2014.Formation of genotoxic compounds by medium pressure ultraviolet treatmentof nitrate-rich water. Desalination and Water Treatment 52 (34e36),6275e6281. https://doi.org/10.1080/19443994.2014.925654.

Mason, T., P�etrier, C., 2004. Ultrasound processes. In: Advanced Oxidation Processesfor Water and Wastewater Treatment. IWA Publishing, London, pp. 185e208.

Mer�enyi, G., Lind, J., Naumov, S., von Sonntag, C., 2010a. Reaction of ozone withhydrogen peroxide (peroxone process): a revision of current mechanistic con-cepts based on thermokinetic and quantum-chemical considerations. Environ.Sci. Technol. 44 (9), 3505e3507. https://doi.org/10.1021/es100277d.

Mer�enyi, G., Lind, J., Naumov, S., von Sonntag, C., 2010b. The reaction of ozone withthe hydroxide ion: mechanistic considerations based on thermokinetic andquantum chemical calculations and the role of HO4- in superoxide dismutation.Chemistry 16 (4), 1372e1377. https://doi.org/10.1002/chem.200802539.

Mitch, W.A., Sharp, J.O., Trussell, R.R., Valentine, R.L., Alvarez-Cohen, L., Sedlak, D.L.,2003. N -nitrosodimethylamine (NDMA) as a drinking water contaminant: areview. Environ. Eng. Sci. 20 (5), 389e404. https://doi.org/10.1089/109287503768335896.

Müller, J.P., Gottschalk, C., Jekel, M., 2001. Comparison of advanced oxidation pro-cesses in flow-through pilot plants (part II). Water Sci. Technol.: J. Int. Assoc.Water Pollut. Res. 44 (5), 311e315.

Nawrocki, J., Kasprzyk-Hordern, B., 2010. The efficiency and mechanisms of catalyticozonation. Appl. Catal. B Environ. 99 (1-2), 27e42. https://doi.org/10.1016/j.apcatb.2010.06.033.

Nickelsen, M., 1994. High energy electron beam generation of oxidants for thetreatment of benzene and toluene in the presence of radical scavengers. WaterRes. 28 (5), 1227e1237. https://doi.org/10.1016/0043-1354(94)90211-9.

N€othe, T., Fahlenkamp, H., von Sonntag, C., 2009. Ozonation of wastewater: rate ofozone consumption and hydroxyl radical yield. Environ. Sci. Technol. 43 (15),5990e5995. https://doi.org/10.1021/es900825f.

Oliveira, H.G., Ferreira, L.H., Bertazzoli, R., Longo, C., 2015. Remediation of 17-a-ethinylestradiol aqueous solution by photocatalysis and electrochemically-assisted photocatalysis using TiO2 and TiO2/WO3 electrodes irradiated by asolar simulator. Water Res. 72, 305e314. https://doi.org/10.1016/j.watres.2014.08.042.

Oppenl€ander, T., 2003. Photochemical Purification of Water and Air: AdvancedOxidation Processes (AOPs): Principles, Reaction Mechanisms, Reactor Con-cepts. Wiley-VCH, Weinheim.

Pang, Y.L., Abdullah, A.Z., Bhatia, S., 2011. Review on sonochemical methods in thepresence of catalysts and chemical additives for treatment of organic pollutantsin wastewater. Desalination 277 (1e3), 1e14. https://doi.org/10.1016/j.desal.2011.04.049.

Parsons, S. (Ed.), 2004. Advanced Oxidation Processes for Water and WastewaterTreatment, Reprinted. IWA Publishing, London.

Pines, D.S., Reckhow, D.A., 2002. Effect of dissolved cobalt(II) on the ozonation ofoxalic acid. Environ. Sci. Technol. 36 (19), 4046e4051. https://doi.org/10.1021/es011230w.

Pisarenko, A.N., Stanford, B.D., Yan, D., Gerrity, D., Snyder, S.A., 2012. Effects of ozoneand ozone/peroxide on trace organic contaminants and NDMA in drinkingwater and water reuse applications. Water Res. 46 (2), 316e326. https://doi.org/10.1016/j.watres.2011.10.021.

Piscitelli, D., Zingaretti, D., Verginelli, I., Gavasci, R., Baciocchi, R., 2015. The fate ofMtBE during Fenton-like treatments through laboratory scale column tests.J. Contam. Hydrol. 183, 99e108. https://doi.org/10.1016/j.jconhyd.2015.10.007.

Qu, X., Alvarez, P.J.J., Li, Q., 2013. Applications of nanotechnology in water andwastewater treatment. Water Res. 47 (12), 3931e3946. https://doi.org/10.1016/j.watres.2012.09.058.

Quan, X., Cheng, Z., Chen, B., Zhu, X., 2013. Electrochemical oxidation of recalcitrantorganic compounds in biologically treated municipal solid waste leachate in aflow reactor. J. Environ. Sci. 25 (10), 2023e2030. https://doi.org/10.1016/S1001-0742(12)60253-8.

Rahim Pouran, S., Abdul Aziz, A.R., Wan Daud, W.M.A., 2015. Review on the mainadvances in photo-Fenton oxidation system for recalcitrant wastewaters. J. Ind.Eng. Chem. 21, 53e69. https://doi.org/10.1016/j.jiec.2014.05.005.

Rahim Pouran, S., Abdul Raman, A.A., Wan Daud, W.M.A., 2014. Review on theapplication of modified iron oxides as heterogeneous catalysts in Fenton re-actions. J. Clean. Prod. 64, 24e35. https://doi.org/10.1016/j.jclepro.2013.09.013.

Rajab, M., Heim, C., Greco, G., Helmreich, B., Letzel, T., 2013. Removal of sulfa-methoxazole from wastewater treatment plant effluents by a boron-dopeddiamond electrode. Int. J. Environ. Pollut. Solut. https://doi.org/10.7726/ijeps.2013.1008.

Rajab, M., Heim, C., Letzel, T., Drewes, J.E., Helmreich, B., 2015. Electrochemicaldisinfection using boron-doped diamond electrode e the synergetic effects ofin situ ozone and free chlorine generation. Chemosphere 121, 47e53. https://doi.org/10.1016/j.chemosphere.2014.10.075.

Reckhow, D.A., Linden, K.G., Kim, J., Shemer, H., Makdissy, G., 2010. Effect of UVtreatment on DBP formation. Am. Water Works Assoc. J. 102 (6), 100.

Redpath, J.L., Willson, R.L., 1975. Chain reactions and radiosensitization: modelenzyme studies. International journal of radiation biology and related studies inphysics. Chem. Med. 27 (4), 389e398. https://doi.org/10.1080/09553007514550361.

Reisz, E., Schmidt, W., Schuchmann, H.-P., Sonntag, C. von, 2003. Photolysis of ozonein aqueous solutions in the presence of tertiary butanol. Environ. Sci. Technol.37 (9), 1941e1948. https://doi.org/10.1021/es0113100.

Richardson, S.D., Plewa, M.J., Wagner, E.D., Schoeny, R., Demarini, D.M., 2007.Occurrence, genotoxicity, and carcinogenicity of regulated and emergingdisinfection by-products in drinking water: a review and roadmap for research.Mutat. Res. 636 (1e3), 178e242. https://doi.org/10.1016/j.mrrev.2007.09.001.

Rosenfeldt, E.J., Linden, K.G., Canonica, S., von Gunten, U., 2006. Comparison of theefficiency of OH radical formation during ozonation and the advanced oxidationprocesses O3/H2O2 and UV/H2O2. Water Res. 40 (20), 3695e3704. https://doi.org/10.1016/j.watres.2006.09.008.

Rosenfeldt, E.J., Melcher, B., Linden, K.G., 2005. UV and UV/H2O2 treatment ofmethylisoborneol (MIB) and geosmin in water. J. Water Supply Res. Technol. -Aqua 54 (7), 423e434.

Safarzadeh-Amiri, A., Bolton, J.R., Cater, S.R., 1996. The use of iron in advancedoxidation processes. J. Adv. Oxid. Technol. 1, 18e26.

S�anchez P�erez, J.A., Rom�an S�anchez, I.M., Carra, I., Cabrera Reina, A., Casas L�opez, J.L.,Malato, S., 2013. Economic evaluation of a combined photo-Fenton/MBR pro-cess using pesticides as model pollutant. Factors affecting costs. J. Hazard Mater.244-245, 195e203. https://doi.org/10.1016/j.jhazmat.2012.11.015.

S�anchez-Polo, M., Gunten, U. von, Rivera-Utrilla, J., 2005. Efficiency of activatedcarbon to transform ozone into *OH radicals: influence of operational param-eters. Water Res. 39 (14), 3189e3198. https://doi.org/10.1016/j.watres.2005.05.026.

Shah, A.D., Dotson, A.D., Linden, K.G., Mitch, W.A., 2011. Impact of UV disinfectioncombined with chlorination/chloramination on the formation of halonitro-methanes and haloacetonitriles in drinking water. Environ. Sci. Technol. 45 (8),3657e3664. https://doi.org/10.1021/es104240v.

Sharpless, C.M., Linden, K.G., 2001. UV photolysis of nitrate: effects of naturalorganic matter and dissolved inorganic carbon and implications for UV waterdisinfection. Environ. Sci. Technol. 35 (14), 2949e2955. https://doi.org/10.1021/es002043l.

Sharpless, C.M., Linden, K.G., 2003. Experimental and model comparisons of low-and medium-pressure Hg lamps for the Direct and H 2 O 2 assisted UV pho-todegradation of N -nitrosodimethylamine in simulated drinking water. Envi-ron. Sci. Technol. 37 (9), 1933e1940. https://doi.org/10.1021/es025814p.

Sichel, C., Garcia, C., Andre, K., 2011. Feasibility studies: UV/chlorine advancedoxidation treatment for the removal of emerging contaminants. Water Res. 45(19), 6371e6380. https://doi.org/10.1016/j.watres.2011.09.025.

Simonsen, M.E., Muff, J., Bennedsen, L.R., Kowalski, K.P., Søgaard, E.G., 2010. Pho-tocatalytic bleaching of p-nitrosodimethylaniline and a comparison to theperformance of other AOP technologies. J. Photochem. Photobiol. Chem. 216(2e3), 244e249. https://doi.org/10.1016/j.jphotochem.2010.07.008.

Song, K., Mohseni, M., Taghipour, F., 2016. Application of ultraviolet light-emittingdiodes (UV-LEDs) for water disinfection: a review. Water Res. 94, 341e349.https://doi.org/10.1016/j.watres.2016.03.003.

Stalter, D., O'Malley, E., von Gunten, U., Escher, B.I., 2016. Fingerprinting the reactivetoxicity pathways of 50 drinking water disinfection by-products. Water Res. 91,



https://doi.org/10.1021/cr00018a003

https://doi.org/10.2166/wst.2011.079

https://doi.org/10.2166/wst.2011.079






https://doi.org/10.1021/ie050981u

https://doi.org/10.1016/j.scitotenv.2013.12.065







https://doi.org/10.1016/j.ultsonch.2009.09.005

https://doi.org/10.1016/j.ultsonch.2009.09.005



https://doi.org/10.1080/19443994.2014.925654





https://doi.org/10.1021/es100277d

https://doi.org/10.1002/chem.200802539

https://doi.org/10.1089/109287503768335896

https://doi.org/10.1089/109287503768335896







https://doi.org/10.1016/0043-1354(94)90211-9

https://doi.org/10.1021/es900825f







https://doi.org/10.1016/j.desal.2011.04.049

https://doi.org/10.1016/j.desal.2011.04.049



https://doi.org/10.1021/es011230w

https://doi.org/10.1021/es011230w



https://doi.org/10.1016/j.jconhyd.2015.10.007



https://doi.org/10.1016/S1001-0742(12)60253-8

https://doi.org/10.1016/S1001-0742(12)60253-8

https://doi.org/10.1016/j.jiec.2014.05.005

https://doi.org/10.1016/j.jclepro.2013.09.013

https://doi.org/10.7726/ijeps.2013.1008

https://doi.org/10.7726/ijeps.2013.1008





https://doi.org/10.1080/09553007514550361

https://doi.org/10.1080/09553007514550361

https://doi.org/10.1021/es0113100

https://doi.org/10.1016/j.mrrev.2007.09.001













https://doi.org/10.1021/es104240v

https://doi.org/10.1021/es002043l

https://doi.org/10.1021/es002043l

https://doi.org/10.1021/es025814p


https://doi.org/10.1016/j.jphotochem.2010.07.008



19e30. https://doi.org/10.1016/j.watres.2015.12.047.Stefan, M.I. (Ed.), 2018. Advanced Oxidation Processes for Water Treatment: Fun-

damentals and Applications. IWA Publishing, London.Stefan, M.I., Bolton, J.R., 2002. UV direct photolysis of N-Nitrosodimethylamine

(NDMA): kinetic and product study. Helv. Chim. Acta 85 (5), 1416. https://doi.org/10.1002/1522-2675(200205)85:5%3c1416:AID-HLCA1416%3e3.0.CO;2-I.

Sun, L., Bolton, J.R., 1996. Determination of the quantum yield for the photochemicalgeneration of hydroxyl radicals in TiO 2 suspensions. J. Phys. Chem. 100 (10),4127e4134. https://doi.org/10.1021/jp9505800.

Sutherland, J., Adams, C., Kekobad, J., 2004. Treatment of MTBE by air stripping,carbon adsorption, and advanced oxidation: technical and economic compari-son for five groundwaters. Water Res. 38 (1), 193e205. https://doi.org/10.1016/j.watres.2003.09.008.

Tr€oster, I., Sch€afer, L., Fryda, M., Matthee, T., 2004. Electrochemical advancedoxidation process using DiaChem electrodes. Water Sci. Technol. 49 (4),207e212.

Ure~na de Vivanco, M., Rajab, M., Heim, C., Letzel, T., Helmreich, B., 2013. Setup andenergetic considerations for three advanced oxidation reactors treating organiccompounds. Chem. Eng. Technol. 36 (2), 355e361. https://doi.org/10.1002/ceat.201200478.

Vallejo, M., Fresnedo San Rom�an, M., Ortiz, I., Irabien, A., 2015. Overview of thePCDD/Fs degradation potential and formation risk in the application ofadvanced oxidation processes (AOPs) to wastewater treatment. Chemosphere118, 44e56. https://doi.org/10.1016/j.chemosphere.2014.05.077.

Vohra, M., Davis, A.P., 2000. TiO2-Assisted photocatalysis of leadeEDTA. Water Res.34 (3), 952e964. https://doi.org/10.1016/S0043-1354(99)00223-7.

von Gunten, U., 2003a. Ozonation of drinking water: Part I. Oxidation kinetics andproduct formation. Water Res. 37 (7), 1443e1467. https://doi.org/10.1016/S0043-1354(02)00457-8.

von Gunten, U., 2003b. Ozonation of drinking water: Part II. Disinfection and by-product formation in presence of bromide, iodide or chlorine. Water Res. 37(7), 1469e1487. https://doi.org/10.1016/S0043-1354(02)00458-X.

von Gunten, U., Oliveras, Y., 1998. Advanced oxidation of bromide-containing wa-ters: bromate Formation mechanisms. Environ. Sci. Technol. 32 (1), 63e70.https://doi.org/10.1021/es970477j.

von Sonntag, C., 2008. Advanced oxidation processes: mechanistic aspects. WaterSci. Technol. 58 (5), 1015e1021. https://doi.org/10.2166/wst.2008.467.

Wacławek, S., Lutze, H.V., Grübel, K., Padil, V.V.T., �Cerník, M., Dionysiou, D.D., 2017.Chemistry of persulfates in water and wastewater treatment: a review. Chem.Eng. J. 330, 44e62. https://doi.org/10.1016/j.cej.2017.07.132.

Wadley, S., Waite, T.D.I., 2004. Fenton process. In: Advanced Oxidation Processes forWater and Wastewater Treatment. IWA Publishing, London, pp. 111e136.

Wang, W.-L., Wu, Q.-Y., Huang, N., Wang, T., Hu, H.-Y., 2016. Synergistic effect be-tween UV and chlorine (UV/chlorine) on the degradation of carbamazepine:influence factors and radical species. Water Res. 98, 190e198. https://doi.org/10.1016/j.watres.2016.04.015.

Wang, W.-L., Wu, Q.-Y., Li, Z.-M., Lu, Y., Du, Y., Wang, T., Huang, N., Hu, H.-Y., 2017.Light-emitting diodes as an emerging UV source for UV/chlorine oxidation:carbamazepine degradation and toxicity changes. Chem. Eng. J. 310, 148e156.https://doi.org/10.1016/j.cej.2016.10.097.

Watts, M.J., Linden, K.G., 2007. Chlorine photolysis and subsequent OH radicalproduction during UV treatment of chlorinated water. Water Res. 41 (13),

2871e2878. https://doi.org/10.1016/j.watres.2007.03.032.Watts, M.J., Rosenfeldt, E.J., Linden, K.G., 2007. Comparative OH radical oxidation

using UV-Cl and UV-HO processes. J. Water Supply Res. Technol. - Aqua 56 (8),469. https://doi.org/10.2166/aqua.2007.028.

Electrochemical wastewater treatment: oxidation of persistent wastewater con-stituents. In: Woisetschl€ager, D., Humpl, B., Koncar, M., Glasl, W.,Siebenhofer, M. (Eds.), 2015. Presentation at the 15th AIChE Annual Meeting.

Wols, B.A., Hofman-Caris, C.H.M., 2012. Review of photochemical reaction constantsof organic micropollutants required for UV advanced oxidation processes inwater. Water Res. 46 (9), 2815e2827. https://doi.org/10.1016/j.watres.2012.03.036.

Wols, B.A., Hofman-Caris, C.H.M., Harmsen, D.J.H., Beerendonk, E.F., 2013. Degra-dation of 40 selected pharmaceuticals by UV/H2O2. Water Res. 47 (15),5876e5888. https://doi.org/10.1016/j.watres.2013.07.008.

Wu, D., Liu, Y., He, H., Zhang, Y., 2016a. Magnetic pyrite cinder as an efficient het-erogeneous ozonation catalyst and synergetic effect of deposited Ce. Chemo-sphere 155, 127e134. https://doi.org/10.1016/j.chemosphere.2016.04.041.

Wu, J., Ma, L., Chen, Y., Cheng, Y., Liu, Y., Zha, X., 2016b. Catalytic ozonation oforganic pollutants from bio-treated dyeing and finishing wastewater usingrecycled waste iron shavings as a catalyst: removal and pathways. Water Res.92, 140e148. https://doi.org/10.1016/j.watres.2016.01.053.

Xiao, Y., Zhang, L., Zhang, W., Lim, K.-Y., Webster, R.D., Lim, T.-T., 2016. Comparativeevaluation of iodoacids removal by UV/persulfate and UV/H2O2 processes.Water Res. 102, 629e639. https://doi.org/10.1016/j.watres.2016.07.004.

Xing, S., Lu, X., Liu, J., Zhu, L., Ma, Z., Wu, Y., 2016. Catalytic ozonation of sulfosali-cylic acid over manganese oxide supported on mesoporous ceria. Chemosphere144, 7e12. https://doi.org/10.1016/j.chemosphere.2015.08.044.

Xylem, 2015. City of Los Angeles Terminal Island Water Reclamation Plant SelectsInnovative Water Reuse Solution to Address Need for Water Security. Pressrelease from June 3, 2015. https://goo.gl/92A5B9. (Accessed 14 February 2018).

Yang, L., Chen, Z., Shen, J., Xu, Z., Liang, H., Tian, J., Ben, Y., Zhai, X., Shi, W., Li, G.,2009. Reinvestigation of the Nitrosamine-formation mechanism during ozon-ation. Environ. Sci. Technol. 43 (14), 5481e5487. https://doi.org/10.1021/es900319f.

Yang, Y., Pignatello, J.J., Ma, J., Mitch, W.A., 2014. Comparison of halide impacts onthe efficiency of contaminant degradation by sulfate and hydroxyl radical-basedadvanced oxidation processes (AOPs). Environ. Sci. Technol. 48 (4), 2344e2351.https://doi.org/10.1021/es404118q.

Zhang, R., Sun, P., Boyer, T.H., Zhao, L., Huang, C.-H., 2015. Degradation of phar-maceuticals and metabolite in synthetic human urine by UV, UV/H2O2, and UV/PDS. Environ. Sci. Technol. 49 (5), 3056e3066. https://doi.org/10.1021/es504799n.

Zhang, Y., Geissen, S.-U., Gal, C., 2008. Carbamazepine and diclofenac: removal inwastewater treatment plants and occurrence in water bodies. Chemosphere 73(8), 1151e1161. https://doi.org/10.1016/j.chemosphere.2008.07.086.

Zhihui, A., Peng, Y., Xiaohua, L., 2005. Degradation of 4-chlorophenol by microwaveirradiation enhanced advanced oxidation processes. Chemosphere 60 (6),824e827. https://doi.org/10.1016/j.chemosphere.2005.04.027.

Zhuo, Q., Deng, S., Yang, B., Huang, J., Yu, G., 2011. Efficient electrochemical oxida-tion of perfluorooctanoate using a Ti/SnO2-Sb-Bi anode. Environ. Sci. Technol.45 (7), 2973e2979. https://doi.org/10.1021/es1024542.




https://doi.org/10.1002/1522-2675(200205)85:5%3c1416:AID-HLCA1416%3e3.0.CO;2-I

https://doi.org/10.1002/1522-2675(200205)85:5%3c1416:AID-HLCA1416%3e3.0.CO;2-I

https://doi.org/10.1021/jp9505800









https://doi.org/10.1002/ceat.201200478

https://doi.org/10.1002/ceat.201200478


https://doi.org/10.1016/S0043-1354(99)00223-7

https://doi.org/10.1016/S0043-1354(02)00457-8

https://doi.org/10.1016/S0043-1354(02)00457-8

https://doi.org/10.1016/S0043-1354(02)00458-X

https://doi.org/10.1021/es970477j

https://doi.org/10.2166/wst.2008.467









https://doi.org/10.2166/aqua.2007.028












https://goo.gl/92A5B9



https://doi.org/10.1021/es404118q

https://doi.org/10.1021/es504799n

https://doi.org/10.1021/es504799n



https://doi.org/10.1021/es1024542

evaluation of advanced oxidation processes for water and

Documents