Sequential Combination of Electro-Fenton and Electrochemical

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Sequential Combination of Electro-Fenton and Electrochemical Chlorination Processes for Treatment of Anaerobically-Digested Food Wastewater Yong-Uk Shin, Ha-Young Yoo, Seonghun Kim, Kyung-Mi Chung, Yong-Gyun Park, Kwang-Hyun Hwang, Seokwon Hong, Hyunwoong Park, Kangwoo Cho, and Jaesang Lee Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b02018 • Publication Date (Web): 09 Aug 2017 Downloaded from http://pubs.acs.org on August 9, 2017

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Environmental Science & Technology

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Sequential Combination of Electro-Fenton and Electrochemical

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Chlorination Processes for Treatment of Anaerobically-Digested

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Food Wastewater

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Yong-Uk Shina,b, Ha-Young Yooa, Seonghun Kimc, Kyung-Mi Chungd, Yong-Gyun Parkd,

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Kwang-Hyun Hwangd, Seok Won Hongb,e, Hyunwoong Parkc, Kangwoo Chof, and Jaesang

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Leea,b*

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a

Civil, Environmental, and Architectural Engineering, Korea University, Seoul 136-701, Korea

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b

Energy Environmental Policy and Technology, Green School, Korea University-KIST, Seoul 136-701,

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Korea

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c

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d

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Korea

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e

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Korea

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f

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(POSTECH), Pohang 790-784, Korea

School of Energy Engineering, Kyungpook National University, Daegu 41566, Korea Environment Process Engineering Team, GS Engineering and Construction Corporation, Seoul 110-789,

Center for Water Resource Cycle Research, Korea Institute of Science and Technology, Seoul 136-791,

Division of Environmental Science and Engineering, Pohang University of Science and Technology

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Submitted to

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*Corresponding author.

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Phone: +82-2-3290-4864; Fax: +82-2-928-7656; E-mail: [email protected]

ACS Paragon Plus Environment


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Abstract

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A two-stage sequential electro-Fenton (E-Fenton) oxidation followed by electrochemical

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chlorination (EC) was demonstrated to concomitantly treat high concentrations of organic carbon

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and ammonium nitrogen (NH4+-N) in real anaerobically-digested food wastewater (ADFW). The

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anodic Fenton process caused the rapid mineralization of phenol as a model substrate through the

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production of hydroxyl radical as the main oxidant. The electrochemical oxidation of NH4+ by a

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dimensionally stable anode (DSA) resulted in temporal concentration profiles of combined and

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free chlorine species that were analogous to those during the conventional breakpoint

30

chlorination of NH4+. Together with the minimal production of nitrate, this confirmed that the

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conversion of NH4+ to nitrogen gas was electrochemically achievable. The monitoring of

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treatment performance with varying key parameters (e.g., current density, H2O2 feeding rate, pH,

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NaCl loading, DSA type) led to the optimization of two component systems. The comparative

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evaluation of two sequentially-combined systems (i.e., the E-Fenton/EC system versus the EC/E-

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Fenton system) using the mixture of phenol and NH4+ under the predetermined optimal

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conditions suggested the superiority of the E-Fenton/EC system in terms of treatment efficiency

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and energy consumption. Finally, the sequential E-Fenton/EC process effectively mineralized

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organic carbon and decomposed NH4+-N in the real ADFW without external supply of NaCl.

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Keywords: anodic Fenton oxidation, electrochemical breakpoint chlorination, digested food

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wastewater, hydroxyl radical, active chlorine species

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INTRODUCTION

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The augmentation of food supply to fulfill the demands of a growing global population

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has caused a substantial increase in the production of food waste1 (nearly 30% to 40% of the

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food supply is wasted in the United States),2 which likely accounts for most of the municipal

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solid waste.3,

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mitigating marine pollution, almost 90 countries currently prohibit the direct dumping of wastes

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into the ocean; this has been criticized as a short-sighted approach to waste management. With

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the increasing yield of food waste, this restriction on waste disposal has encouraged research

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activity to explore alternative treatment technologies (e.g., pyrolysis5 and aerobic composting6)

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for tackling the issue of food waste. Among the feasible remediation strategies,7 anaerobic

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digestion has been broadly applied for the treatment of food waste due to the high content of

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biologically-degradable organic substrates.1,

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derived organics achieves concomitant removal of organic matter and the generation of biogas

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(i.e., methane). However, the effluent released after the biogasification still contains a high

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concentration of non-biodegradable residual carbon (typically 3000 – 6000 mg/L as COD).10, 11

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Furthermore, the anaerobic degradation of proteins produces a high level of ammonium-nitrogen

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(NH4+-N) in the effluent,10, 11 and excess NaCl is also detected due to the salt-rich diet.11 Such

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chemical composition likely causes the anaerobically-digested food wastewater (ADFW) to

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resist or inhibit the biological treatment processes.12-14 An appropriate technical option therefore

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needs to be developed for ADFW remediation.

4

In compliance with the London Convention that was established in 1972 for

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The microbial transformation of food waste-

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Electrochemical water treatment processes have attracted increasing interest over recent

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years due to their following advantages: 1) electrically-controllable operation under ambient

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conditions, 2) decreased usage of chemical additives, and 3) facile integration with other



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treatment processes or with renewable power generation systems (e.g., solar cell).15,

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Furthermore, electrochemical redox reactions involve the high-yield production of various

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oxidants such as ozone,17 hydroxyl radical (•OH),18 and hypochlorous acid/hypochlorite

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(HOCl/OCl−)19; this allows the application of electrochemical processes for the oxidative

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degradation of a wide-range of aquatic pollutants. For example, the electrochemical generation

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of non-selective •OH directly from water at some anodes (e.g., boron-doped diamond (BDD),

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PbO2) led to the treatment of recalcitrant wastewaters and leachates.16, 18, 20 Powerful oxidizing

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capacity of BDD enabled the electrochemical degradation of some pollutants that are unreactive

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toward •OH (e.g., cyanuric acid,21 perfluorinated compounds22), but likely caused the formation

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of toxic oxyanion byproducts such as bromate and perchlorate in the presence of excess halide

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ions.23 Alternatively, chemical reagents (e.g., Fe2+, H2O2) released or formed electrochemically

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also mediate •OH-induced mineralization and the associated reduction of the chemical oxygen

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demand (COD).24, 25 The electrolysis of chloride-containing electrolytes (e.g., NaCl, KCl) using

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RuO2- or IrO2-based dimensionally stable anodes (DSAs) yields active chlorine species, which

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are also capable of the rapid decomposition of non-biodegradable organics.19 In particular, the

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electrochemical oxidation of ammonia/ammonium ion (NH3/NH4+) at the DSAs proceeds

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through the breakpoint chlorination in which HOCl forms electrochemically and subsequently

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chlorinates NH3 or NH4+ into dinitrogen with minimal yield of nitrite/nitrate (NO2−/NO3−).26

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Based on the aforementioned capability to mineralize organics into CO2 and convert

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NH3/NH4+ into N2, the two-stage electrochemical system consisting of the electro-Fenton (E-

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Fenton; anodic Fenton) and electrochemical chlorination (EC) processes has the potential to be a

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solution that aligns with the technical requirement, i.e. treatability of high-levels of COD and

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NH4+-N in the ADFW. The E-Fenton process has been demonstrated to be highly suitable for


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COD removal in diverse real wastewaters with high organic contents.24, 25 In the anodic Fenton

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process, the release of Fe2+ at the sacrificial iron anode that is accompanied by Fe2+ regeneration

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and H2O2 formation at the cathode enables high efficiency of •OH production with minimal

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formation of iron sludge. While the •OH-generating systems fully oxidize NH3 to NO3− in

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alkaline conditions,27 both NH3 and NH4+ are selectively transformed to N2 via the breakpoint

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chlorination mechanism, which is electrochemically achievable with DSAs.26 Hence, the

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integration with the EC process would contribute mainly to the removal of NH3/NH4+ in the

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ADFW. The electrochemically-generated active chlorine species also oxidize organic

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compounds,28 but the substrate-specific oxidizing capacity does not allow the complete

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mineralization of organics. The BDD-based electrochemical systems can also achieve the abiotic

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mineralization of organic carbon and conversion of NH4+-N into nitrogen gas in the

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wastewaters.29 However, the choice of the combined E-Fenton/EC process should mitigate risk

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resulting from toxic oxyanion formation (note that the formation of chlorate and perchlorate

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more readily occurs on highly oxidizing BDD anodes compared to metal oxide-based anodes23).

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This study explores the possible application of the coupled E-Fenton/EC system for

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treatment of the ADFW containing refractory organics and NH4+-N at high concentrations. The

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sequential combinations employing electrochemical oxidation processes (e.g., E-Fenton/photo-

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Fenton,30 E-Fenton/anaerobic digestion,31 E-Fenton/photoelectrocatalysis32) have been proposed

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in the literature, but the strategies aimed to synergistically improve the overall efficiency of

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organic carbon removal in most cases. On the other hand, this study examines the combined E-

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Fenton/EC process for concomitant treatment of non-biodegradable organic carbon and NH4+-N

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in the high-salinity wastewater. The main mechanism of the E-Fenton process is investigated

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based on the hydroxylation efficiency of benzoic acid, the quenching effect of methanol as a



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radical scavenger, and multi-activity assessment. The total nitrogen loss and time-dependent

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concentration variation of combined and free chlorines confirm the possible occurrence of the

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breakpoint chlorination at the DSA. The effects of the operation parameters including the initial

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pH, current density, H2O2 feeding rate, type of DSA, and electrolyte concentration on the

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performance of the E-Fenton and EC processes are investigated. Two integrated systems that

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couple the E-Fenton process with the EC process (i.e., the E-Fenton/EC system versus the EC/E-

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Fenton system) are comparatively evaluated in terms of the treatment efficiency of the binary

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mixture of phenol and NH4+. Finally, the E-Fenton process in combination with the EC process

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is examined under the optimized operating conditions for the concomitant treatment of organic

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carbon and NH4+-N in a real ADFW.

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MATERIALS AND METHODS

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Chemicals and Materials. All chemicals were of reagent grade and were used without further

125

purification or treatment (see Supporting Information) except for 2,4-dinitrophenyl hydrazine

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(DNPH), which was recrystalized with acetonitrile three times prior to use. Ultrapure deionized

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water (>18 MΩ•cm), produced using a Millipore system, was used for the preparation of all

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experimental solutions.

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Experimental Procedure and Analytical Methods. The electrochemical experiments were

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performed in a magnetically-stirred one-compartment reactor having a volume of 1 L (E-Fenton

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process) or 0.3 L (EC process). The electrolysis current was supplied using a direct current (DC)

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power supply (TPM-1010, Toyotech Co. Korea). The anodes and cathodes were rectangular

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plates with an active geometric area of 50 cm2 (E-Fenton process) or 12 cm2 (EC process), and

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were positioned in parallel with an inter-electrode gap of 1 cm. Since the reactors for the E-


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Fenton and EC processes have solution capacities of 0.7 L and 0.125 L, the specific surface areas

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(as; surface area to volume ratios) of iron anode and DSA were determined to be 7.1 m-1 and 9.6

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m-1, respectively. NaCl at a concentration of 1 M was typically used as a supporting electrolyte.

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In the E-Fenton process, ferrous ion was transferred to the undivided electrolytic reactor through

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the electrolysis at a sacrificial stainless steel anode, and 5.83 M H2O2 stock solution was

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continuously delivered at a constant feeding rate (typically 1.67 mmol/min) by a peristaltic pump

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(Minipuls3 Gilson co, France). A stainless steel plate of the same size was used as the cathode.

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The EC process was performed using commercially available DSA (Nanopac DSA (NP-DSA);

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Samsung DSA (S-DSA); Ko-Electrode DSA (KE-DSA)) as an anode and a Ti plate (Aldrich) as

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a cathode. The SEM images of the three DSAs confirm the presence of crystallite agglomerates

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and dried mud cracks on the flat electrodes as the morphological features typical of DSA

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(Figures S1-S3).

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To monitor the efficiency of the electrochemical oxidation of organic pollutants (e.g.,

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phenol, nitrobenzene), aliquots of 1 mL were withdrawn from the electrolysis reactor at pre-

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determined time intervals using a 1-mL syringe. The aliquots were then filtered through a 0.45-

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µm PTFE filter (Whatman), and were transferred into a 2-mL amber glass vial for further

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analysis of target substrates. The concentration of organic compounds was measured using a

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high-performance liquid chromatography (HPLC, Agilent Infinity 1260) system equipped with a

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C-18 column (ZORBAX Eclipse XDB-C18) and a UV/Vis detector (G1314F 1260VWD). The

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HPLC analysis was carried out with the eluent consisting of a binary mixture of 0.1% (v/v)

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aqueous phosphoric acid and neat acetonitrile. Dissolved organic carbon was monitored by a

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total organic carbon analyzer (TOC-VCPH, Shimadzu). Total dissolved nitrogen (TDN) was

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monitored using Humas HS-TN-L (CA) reagent during the electrochemical ammonia oxidation;



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aqueous nitrogen-containing compounds are converted into NO3− through the alkaline persulfate

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digestion, which is followed by the colorimetric determination of NO3− based on the

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chromotropic acid method.33 The quantification of NH4+, NO2−, and NO3− was performed using

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an ion chromatograph (IC, Dionex DX-120). The IC system was equipped with a Dionex IonPac

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AS-14 for detection of NO2− and NO3−, a Dionex IonPac CS-12A for detection of NH4+, and a

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conductivity detector. To explore the possible formation of N-nitrosodimethylamine (NDMA) as

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the toxic disinfection byproduct, NDMA was quantified during the electrochemical chlorination

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of aqueous solutions containing ammonia only, the binary mixture of ammonia and phenol, and

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the ternary mixture of ammonia, phenol, and dimethylamine (DMA). NDMA measurement was

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performed by the HPLC (Ultimate 3000, Dionex) with post-column UV photolysis/Griess

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reaction34; a post-column UV254 irradiation reactor (LCTech Ltd.) coupled with the Griess

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reaction (installed between the column and the photo-diode array detector) photochemically

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transforms nitrosamines into NO2− and subsequently detects the photo-produced NO2− according

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to the colorimetric Griess reaction. Qualitative identification of gaseous nitrogen-based products

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such as N2, NO, N2O, NO2, N2O5, and volatile chloramines was carried out using a quadrupole

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mass spectrometer (QMS; HPR20, Hiden Analytical).

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employed as the source of ammonia in order to distinguish atmospheric nitrogen from nitrogen

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formed via breakpoint chlorination of ammonia. The gas trapped in the reactor headspace was

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transferred to a sealed plastic bag and was analyzed by injection into a QMS. The amount of

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H2O2 remaining in the reactor was determined by the spectrophotometric method using DMP and

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copper (II) (i.e., ε454 = 14300 M-1 cm-1),35 and the electrolytic production of Fe2+ was monitored

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by the colorimetric assay using 1,10-phenanthroline (i.e., ε515 = 11050 M-1 cm-1).36 The

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concentrations of free chlorine and total chlorine were quantitatively analyzed using the

15

NH4Cl (98% atom


15

N, Aldrich) was

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colorimetric method based on the DPD (N,N-diethyl-p-phenylenediamine) chemistry (utilizing

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Humas colorimeter test kits (HS10020-10; HS10030-10)).37 The concentration of combined

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chlorine (i.e., monochloramine (NH2Cl) and dichloramine (NHCl2)) was determined by

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computing the difference between the concentrations of total chlorine and free chlorine. NH2Cl

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was monitored using the Hach MonochlorF reagent as the mixture of sodium nitroferricyanide,

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sodium hydroxide, and phenol.38 In this method, the reaction between NH2Cl and phenol (or

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substituted phenols) catalyzed by cyanoferrates leads to the formation of benzoquinone

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monoamine compounds. The intermediates undergo coupling reactions in the presence of excess

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substituted phenols to form green-colored indophenols, which can be quantified by monitoring

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the absorbance at 655 nm. Formaldehyde as the product of electrochemical methanol oxidation

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was measured using the HPLC after the derivatization in the acidified DNPH solution.39

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Electrochemical Measurement. To estimate the potentials of stainless steel, DSA, and Ti

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electrodes, the open circuit potentials (Eoc) of the electrodes were recorded using a potentiostat

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(Versastat, Princeton) in the three-electrode system with the saturated calomel electrode (SCE)

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as a reference and platinum wire as a counter electrode. At the same time, the potential

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differences between the anodes and cathodes were monitored in the E-Fenton and EC processes

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using a digital multimeter (Agilent 34401A). The Eoc values of the stainless steel electrodes that

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served as the anode and cathode (i.e., anodic and cathodic potentials) were fixed at 0.0 V and –

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1.1 V, respectively (Figure S4a). The potential difference between two iron plate electrodes was

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maintained constantly at 1.1 V during the electrolysis, which was in accord with the ∆Eoc of the

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steel electrodes (i.e., 0.0 V – (–1.1 V)). The Eoc values of DSA and Ti electrodes in the EC

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process were determined to be +2.5 V and –2.5 V, respectively, and the experimentally-

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measured potential difference between two electrodes (5.0 V) was also similar to the ∆Eoc



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(Figure S4b). To evaluate the possibility for some redox processes to take place, the measured

205

potentials with respect to SCE (VSCE) were converted into NHE (normal hydrogen electrode)

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scale (VNHE) according to the equation: VNHE = VSCE + 0.241 – 0.059 × pH.

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Wastewater Characterization. The ADFW was provided by the GS Engineering and

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Construction Corporation’s food waste treatment plant located in Environmental Energy Town

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(Daegu, Korea). The ADFW was subjected to centrifugal separation (3000 rpm for 20 min) using

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a centrifugal dehydrator (SHT350MCH, SHT, Korea) prior to use in the electrochemical

211

experiments. The supernatant was stored in a refrigerator set at 4 oC. The quality parameters of

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the ADFW are listed in Table 1.

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RESULTS AND DISCUSSION

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Oxidative Degradation of Phenol by the E-Fenton Process. Figure 1 shows that the E-Fenton

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process effectively decomposed and mineralized phenol as a model organic substrate. The

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complete degradation of 10 mM phenol was achieved within 40 min when the constant current

218

density of 28.5 mA/cm2 was applied. Further treatment with the E-Fenton oxidation caused a

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TOC reduction of approximately 75%. The experimental solutions were adjusted to pH 3.0, and

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the continuous monitoring of pH and the periodic injection of HClO4 maintained a constant pH

221

over the course of the E-Fenton oxidation. H2O2 externally supplied at a feeding rate of 1.67

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mmol/min was gradually accumulated without electric current, and the electrolysis of an iron

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plate with the dimensions of 5 cm × 10 cm in the absence of H2O2 led to Fe2+ production at a rate

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of 1.13 mmol/min (Figure 1a). Alternatively, the steady-state concentrations of H2O2 and Fe2+

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did not exceed ca. 6% of continuously-injected H2O2 and electrochemically-generated Fe2+

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during the E-Fenton oxidation of phenol (Figure 1a). This implies the rapid reaction of


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externally-injected H2O2 with electrochemically-released Fe2+ from the iron sacrificial anode

228

(leading to •OH production). The comparison of the Eoc values of iron electrodes (anodic

229

potential = +0.064 VNHE; cathodic potential = –0.936 VNHE) with H2 and O2 evolution potentials

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at pH 3 (E(H+/H2) = –0.177 VNHE40; E(O2/H2O) = +1.053 VNHE40) implies that water electrolysis

231

could produce H2 whereas O2 production was not thermodynamically favored. Considering the

232

standard redox potentials of •OH, Cl•, and Cl2 (E0(•OH/H2O) = +2.73 VNHE41; E0(•Cl/Cl–) =

233

+2.41 VNHE41; E0(Cl2/Cl–) = +1.36 VNHE40), we can out the possible contribution of the reactive

234

intermediates that might form during the electrolysis of aqueous chloride solutions on the iron

235

anodes.

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The oxidative degradation of phenol was achievable even with a relatively insufficient

237

supply of Fe2+ and H2O2 (i.e., with a 2 cm × 10 cm iron plate and H2O2 mass input rate of 0.05

238

mmol/min), and the kinetic rates were gradually accelerated with increasing current density

239

(Figure S5). However, such oxidizing capacity of the system is likely attributable to the direct

240

electron transfer at the iron anode based on the insignificant TOC loss (Figure S5) and the

241

negligible benzoic acid degradation (note that the hydroxylation of benzoic acid indirectly

242

indicates •OH formation42) (Figure S6). On the other hand, when substantially increasing the

243

amounts of Fe2+ and H2O2 injected into the E-Fenton system, 4-hydroxylbenzoic acid was

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formed as a hydroxylation product and subsequently underwent further degradation (Figure 1b),

245

which confirms the involvement of •OH in the oxidative treatment in the E-Fenton process. The

246

performance of the E-Fenton process in the oxidative conversion of methanol (as an indicator of

247

•OH

248

effectively produced from methanol when applying high feeding rates of Fe2+ and H2O2 (Figure

249

1b), whereas formaldehyde yield was relatively low in the E-Fenton process where only direct

42

) to formaldehyde also depended on the choice of operation parameters; formaldehyde was



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250

electron transfer is allowed (Figure S6). The switching of the reaction pathway (i.e., direct

251

electron transfer versus radical-induced oxidation) in response to the Fe2+ and H2O2 feeding rates

252

was also observed in the multi-activity assessment using multiple organic pollutants. The E-

253

Fenton process operated at slow feeding rates of Fe2+ and H2O2 showed the substrate-specific

254

reactivity (Figure S7a), which is in marked contrast to the non-selective nature of •OH43;

255

organics such as bisphenol A, 4-chlorophenol, and carbamazepine were effectively oxidized,

256

whereas 4-nitrobenzene and benzoic acid were marginally decomposed. On the other hand, the

257

increased input rates of Fenton reagents caused effective degradation of all tested organic

258

substrates (Figure S7b), which suggests that •OH serves as a main oxidant in the E-Fenton

259

process.

260

Effects of Current Density and H2O2 Feeding Rate on the E-Fenton Oxidation Efficiency.

261

The E-Fenton oxidation of phenol was performed with increasing current density and H2O2

262

feeding rate (Figures 2). The increase in the accumulated concentrations of Fe2+ and H2O2 led to

263

the gradual acceleration in phenol oxidation, which is likely attributed to the enhanced

264

production of •OH at the high concentration of Fenton reagent. The kinetic rate of the E-Fenton

265

oxidation reached saturation point when the current density and H2O2 input rate exceeded 28.5

266

mA/cm2 and 1.67 mmol/min, respectively. No further improvement in the efficiency of the E-

267

Fenton process was observed from the possible competition between Fenton reagent versus

268

organic substrate for •OH,44 which becomes more pronounced in the presence of excess Fe2+ and

269

H2O2.

270

Electrochemical Chlorination of Ammonium Ions at the DSAs. Electrochemical oxidation of

271

NH4+ was performed at the NP-DSA that enables effective production of active chorine species

272

in the presence of chloride-based electrolytes (e.g., NaCl and KCl) (Figures 3a and S8).19, 45 The


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kinetic rate of NH4+ oxidation was continuously enhanced until the applied current density

274

increased to 250 mA/cm2 (further increase in the current density was not allowed due to the

275

capacity of power supply). Electrolysis in the presence of Na2SO4 caused significant deceleration

276

in NH4+ decay, whereas NH4+ oxidation was not kinetically retarded with KCl as the alternative

277

chloride salt (Figure S8). Furthermore, NH4+ oxidation efficiency increased in direct proportion

278

to NaCl concentration up to 1 M (Figure S8). The results reveal the involvement of active

279

chlorine species (i.e., HOCl/OCl−) in the electrochemical decomposition of NH4+. The anodic

280

potential of NP-DSA (i.e., +2.45 VNHE) is sufficiently positive for production of active chlorine

281

species such as •Cl and Cl2 whereas •OH production is not readily achievable. The potentials of

282

DSA and Ti electrodes (cathodic potential = –2.55 VNHE) allow H2 and O2 evolution from water

283

in the EC process. O2 is a mild oxidant compared to active chlorine species and the abundant

284

supply of O2 is expected since all experiments were performed under air-equilibrated conditions,

285

which collectively suggests a minor role of O2 in the EC process.

286

Figure 3b shows that the concentration of combined chlorine gradually increased and

287

reached a peak after 50 min of electrolysis, which results from the reaction between NH4+ and

288

the electrochemically-generated active chlorine (i.e., HOCl dominant at pH 5).46 Time-dependent

289

concertation profiles of NH2Cl and NHCl2 (inset of Figure S10) reveal that NHCl2 was the

290

predominant form of chloramine, and the mass percent of NH2Cl ranged from ca. 2.75% to 34.1%

291

of combined chlorine. The combined chlorine concentration decreased after reaching maximum,

292

which implies that the combined chlorine was transformed to dinitrogen through the breakpoint

293

chlorination mechanism (Reactions 1 and 2).46

294

2NH2Cl + HOCl → N2 + 3HCl + H2O

(1)

295

NH2Cl + NHCl2 → N2 + 3HCl

(2)



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The effective consumption of active chlorine species during the chlorination of NH4+ and

297

chloramine rendered free chlorine undetectable after the initial 90 min, but the apparent

298

production of HOCl drastically increased as the level of NH4+ significantly declined (breakpoint)

299

(Figure 3b). Such temporal concentration profiles of combined and free chlorine species were

300

observed in the conventional breakpoint chlorination of NH3/NH4+,46, 47 which confirms that the

301

oxidative conversion of NH3/NH4+ to nitrogen gas was electrochemically achieved with DSA.

302

NO3− as an undesirable product was formed during the electrochemical degradation of

303

NH4+, and the saturation concentration corresponds to only ca. 2% to 4% of NH4+ initially added

304

in most cases (Figure S9), which accords with the previously-reported ratio of NO3− formed to

305

NH4+ degraded (ca. 2.2%48). NO3− is known to be mediated by dichloramine in the breakpoint

306

chlorination regime (Reaction 3).49

307

NHCl2 + 2HOCl + H2O → NO3− + 4HCl + H+

(3)

308

Note that the NO3− formation yield during the chemical chlorination (i.e., adding NaOCl) of

309

ammonia ranges from 1.5% to 10%.47 Relatively low yields of NO3− in the EC process are

310

ascribed to the formation of active chlorine at a very low concentration in the pseudo steady-state

311

until breakpoint (Figure 3b). A significant TDN reduction was observed during electrochemical

312

NH4+ oxidation, and the measured TDN was in good agreement with the sum of NH4+, NO2−,

313

NO3−, and combined chlorines (Figure S10). The result implies that NH4+ initially injected was

314

predominantly transformed into gaseous nitrogen-containing products. The undetectable amount

315

of toxic NDMA (detection limit < 0.1 µM) was produced when pure NH4+ solution and binary

316

mixture consisting of NH4+ and phenol were electrochemically treated (Figures S11a and S11b).

317

NDMA was found in the concentration range of 0.17 µM to 0.69 µM even when initially adding

318

1 mM of DMA as a well-known NDMA precursor to the mixture (Figure S12), and was


Page 15 of 36


319

eventually degraded to an undetectable level after 2 hr of electrolysis. To qualitatively identify

320

the major gas products, we repeated the EC process with 15NH4Cl (98% atom 15N, Aldrich) and

321

performed headspace gas analysis using a QMS. The mass spectra in Figure S13 show the

322

appearance of one major peak at mass number 30, which assures that the conversion of NH4+ to

323

N2 was primarily responsible for the significant TDN loss and rules out the possibility for

324

nitrogen oxides (e.g., NO, NO2) and chloramines to form as the main gaseous products.

325

Effects of DSA Composition and Initial pH on NH3/NH4+ Oxidation Efficiency. Figure 4a

326

shows a comparison of the efficiency of three commercial DSAs (NP-DSA, KE-DSA, and S-

327

DSA) for the electrochemical oxidation of NH4+ under variable current density conditions. While

328

NP-DSA and KE-DSA exhibited comparable performance in NH4+ decomposition over a wide

329

range of current densities, the XRD analysis of the two electrodes suggested a difference in

330

surface chemical composition (Figure S1 versus Figure S3); IrO2 was mainly coated on the KE-

331

DSA, while IrO2 and TiO2 co-existed on the NP-DSA. Although the heterojunction doping of

332

TiO2 on IrO2 was reported to favor active chlorine generation and prevent IrO2 leaching,50 the

333

enhancing effect of TiO2 was not pronounced. Theoretical and experimental evidence has

334

suggested that the electrocatalytic activity for chlorine evolution increases in the following order:

335

RuO2 > IrO2 >> TiO2.51 Accordingly, the S-DSA surface-modified with IrO2 and RuO2 as

336

electrocatalytically-active components likely showed a moderately-enhanced capacity for NH4+

337

oxidation with current density values lower than 100 mA/cm2 (Figure 4a). On the other hand,

338

regardless of the DSA selected, no significant difference was observed in the NH4+ removal rate

339

at the current density exceeding 100 mA/cm2 (Figures 4a and S14). Under such high current

340

density conditions, the heterogeneous generation of active chlorine might overwhelm the side

341

reactions (e.g., oxygen evolution reaction), irrespective of the surface composition of DSA.



Page 16 of 36

342

The kinetic rate of electrochemical NH3/NH4+ oxidation at the NP-DSA was monitored

343

under variable current density and initial pH conditions (Figure 4b). NH3/NH4+ degradation

344

efficiency varied in direct proportions to the applied current density, regardless of the initial pH.

345

On the other hand, the electrocatalytic activity of NP-DSA for NH3/NH4+ decomposition appears

346

to be dependent on the initial pH; weak acidic and neutral pH conditions (i.e., pH 5 and pH 7,

347

respectively) favored the electrochemical oxidation of NH3/NH4+ over a wide range of current

348

densities, while the rate of NH4+ decomposition was drastically retarded by adjusting the initial

349

pH to 3. The decreased efficiency at pH 3 results from the pH-dependent hydrolysis of

350

electrochemically-generated chlorine gas (Cl2) (Cl2 + H2O ↔ HOCl + H+ + Cl−); Cl2 as the

351

predominant chlorine species at pH 3 undergoes evaporation from the solutions, which leads to

352

significant chlorine loss.52 Furthermore, the reactivity of OCl− (dominant active chlorine at pH >

353

7.5) and the dissolved Cl2 toward NH4+/NH3 is insignificant,53 which explains the maximal

354

efficiency for ammonia removal at pH 5 to 7.

355

Performance Comparison of the E-Fenton/EC System versus the EC/E-Fenton System. We

356

comparatively evaluated two sequentially coupled systems, E-Fenton oxidation followed by EC

357

(i.e., the E-Fenton/EC system) versus EC followed by E-Fenton oxidation (i.e., the EC/E-Fenton

358

system) based on the treatment efficiency of the binary aqueous mixture of phenol and NH4+ as

359

synthetic ADFW. When applying the E-Fenton/EC system under the predetermined optimal

360

condition (i.e., current density = 28.5 mA/cm2; H2O2 feeding rate = 1.67 mmol/min), the E-

361

Fenton oxidation as a pre-treatment process caused a marked reduction in the TOC level, while

362

removing ca. 25% of initial NH4+ (Figure 5a). It is worth noting that phenol mineralization was

363

marginally retarded when phenol and NH4+ co-existed, with kphenol = 0.0122 ± 0.007 min-1 in the

364

absence of NH4+ versus kphenol = 0.0156 ± 0.001 min-1 in the presence of NH4+ (Figure 1a


Page 17 of 36


365

compared to Figure 5a). This results from the low susceptibility of NH4+ to •OH-induced

366

oxidation,27 which generated no significant competition between phenol and NH4+ for •OH in the

367

acidified E-Fenton system. Furthermore, such ineffective oxidation of NH4+ is environmentally

368

desirable considering that NO3− is the main oxidation product when water contaminated with

369

NH3/NH4+ is subjected to advanced oxidation processes (AOPs).27

370

The effluent from the E-Fenton process was delivered to the EC process after adjusting

371

to pH 5, which favored breakpoint chlorination of NH4+ (Figure 4b). Figure 5b shows that NH4+

372

oxidation proceeded more rapidly as the electric current used was increased. No kinetic

373

retardation in the post-process might imply that residual organics, acting as the possible quencher

374

of active chlorine, were present at low concentrations. Furthermore, ammonia is considerably

375

more vulnerable to HOCl-induced oxidation than phenolic compounds, with k(HOCl + NH3) =

376

4.2 × 106 M-1s-1 versus k(HOCl + phenolate) = 3.5 × 104 M-1s-1.53 With the applied current

377

density exceeding 150 mA/cm2, NH4+ was completely decomposed with the production of NO3−,

378

which corresponds to ca. 12% of the amount of initial ammonia (Figure S15). The increased

379

yield of NO3− formation may be attributed to the possible reactions between NH2Cl and Fe2+54 or

380

phenol and its hydroxylated intermediates,55 which could interfere with the reaction route for N2

381

formation (Reaction 1); selective reduction of NH2Cl by Fe2+ or phenolic compounds could

382

kinetically retard the further oxidation of NH2Cl to N2 (Reaction 1) and concurrently increase the

383

availability of NHCl2 for HOCl, which likely favors NO3− production (Reaction 2).

384

In the EC/E-Fenton system, the pre-treatment process (i.e., EC process) allowed rapid

385

NH4+ degradation when applying the current density above 150 mA/cm2 (Figure 5c), which led

386

to NO3− formation with a yield of ca. 0.1 (Figure S16). The NH4+ oxidation efficiency was

387

reduced to a certain extent, which is likely ascribed to the presence of background organics (i.e.,



Page 18 of 36

388

phenol and its oxidation products) that are also vulnerable to chlorination.19 The EC system

389

performed at the highest current density of 200 mA/cm2 was not effective for phenol

390

mineralization, which confirms the superiority of the E-Fenton process in terms of TOC removal.

391

The subsequent E-Fenton oxidation caused the oxidative degradation of residual organics in the

392

effluent, but the rate of mineralization was two-fold slower than when the E-Fenton process was

393

applied prior to the EC process (Figure 5a compared to Figure 5d). The retarded mineralization

394

may be attributable to the formation of chlorinated intermediates/products during the EC process,

395

but clarification of this requires further study.

396

We also comparatively evaluated the E-Fenton process versus the EC process in terms of

397

the major operating cost of mineralization based on the specific energy consumption (SEC) (i.e.,

398

the amount of electric energy required to remove a unit mass of TOC) computed by equation

399

(4).56 SEC =

V × I × t 4 TOC − TOC × V

400

where Vc is the voltage measured during the electrolysis, I is the applied electric current, t is the

401

reaction time, TOC0 and TOCf are the initial and final TOC values, respectively, and V is the

402

volume of the experimental solution. The contribution of the H2O2 injection to the energy

403

consumed to reduce TOC in the E-Fenton process was considered with the H2O2 energy, which

404

is calculated as the ratio of electricity cost to the H2O2 cost (the average ratio was found to be ca.

405

10.81 kWh•kg-1H2O2 for the five local regions surveyed).57 The SEC values for the E-Fenton and

406

EC processes were determined to be 179.30 kWh•kgTOC-1 and 613.20 kWh•kgTOC-1, respectively.

407

Overall, the results presented in Figures 5a-5d suggest that the E-Fenton/EC system outperforms

408

the EC/E-Fenton system in concomitantly treating the mixture of organics and NH4+-N in terms

409

of treatment efficiency and energy consumption.


Page 19 of 36


410

Application of the E-Fenton/EC System for Treatment of the Real DFW. The sequential E-

411

Fenton/EC system was performed to concurrently treat high concentrations of organic carbon (ca.

412

1800 mg/L – 2000 mg/L) and NH4+-N (ca. 3800 mg/L – 4000 mg/L) in a real ADFW matrix

413

under the predetermined optimal operating conditions. While the E-Fenton process as a pre-

414

treatment step enabled significant TOC loss, it decomposed ca. 25% of NH4+ that initially

415

existed in the wastewater matrix (Figure 6a), which confirms the previous finding that the E-

416

Fenton process caused effective •OH-mediated mineralization of organics (Figure 5a). Almost

417

complete NH4+ degradation was achieved when the EC process using the NP-DSA was

418

subsequently applied. NO3− that accounts for ca. 12% of the initial NH4+ and NO2− at an

419

undetectable level was formed as the oxidation products (Figure 6a), which implies that NH4+ in

420

the ADFW was converted predominantly into dinitrogen gas in the post-treatment process. It is

421

worth noting that no external supply of chloride was required to initiate the electrochemical

422

breakpoint chlorination since ADFW inherently contains a sufficient amount of chloride ions.

423

The SEC values for organic carbon mineralization (E-Fenton process) and ammonia removal

424

(EC process) from the ADFW were determined to be 106.10 kWh•kgTOC-1 and 106.30

425

kWh•kgNH4+-1, respectively (the amount of electric energy required to remove a unit mass of

426

NH4+ indicates the SEC for ammonia removal).

427

In an effort to explore the possible improvement in treatment performance with

428

increasing electrochemically-active sites, the coupled E-Fenton/EC process was repeated for

429

ADFW treatment when double pairs of anodes and cathodes were alternatively used in both unit

430

treatment steps. Figure 6b shows that the increase in the number of anode-cathode pairs

431

accelerated organic carbon mineralization and ammonia oxidation to a certain extent, with

432

k(double)/k(single) (the ratio of the pseudo-first order constant obtained with double pairs of



Page 20 of 36

433

anodes and cathodes to that obtained with a single pair) for TOC removal = 2.01 ± 1.20 and

434

k(double)/k(single) for NH4+ oxidation = 1.42 ± 0.87). This result is likely attributable to the

435

enhanced production of oxidizing species including •OH and HOCl/OCl−, which was confirmed

436

based on the increased efficiency of benzoic acid hydroxylation and of HOCl/ OCl− production

437

(Figure S17). However, further studies are needed (e.g., monitoring of treatment efficiency as a

438

function of the electrode number in undivided electrolytic cells consisting of an alternating

439

succession of anodes and cathodes) to confirm the technical advantage of multiple electrode

440

systems.

441

Environmental Application. As a technical strategy to overcome the challenge associated with

442

anaerobically-digested food wastewater effluent, we proposed the sequential integration of two

443

processes, anodic Fenton oxidation and electrochemical chlorination, using DSA. Even if the

444

conventional AOPs can treat non-biodegradable organic carbon via powerful oxidizing radical,

445

•OH,

446

exceeds the pKa of ammonia (i.e., 9.3) due to the very low reactivity of •OH toward NH4+.

447

Furthermore, the non-selective nature of •OH causes the stoichiometric conversion of ammonia

448

to nitrate, which rather increases the toxicity level of wastewater. The shortcoming of AOPs in

449

ammonia treatment was resolved by sequentially coupling E-Fenton process with EC process;

450

electrochemical breakpoint chlorination transformed NH4+ to dinitrogen gas with minimal

451

release of nitrate in the circumneutral pH region. The alternative use of biological method as a

452

post-treatment step possibly leads to the oxidative conversion of NH3/NH4+ to nitrogen gas.

453

However, biological ammonia removal usually proceeds via two-step nitrification-denitrification,

454

and the treatment efficiency is highly vulnerable to temperature, pH, ammonia concentration,

455

and salinity. In contrast, electrochemical oxidation on DSA converts NH3/NH4+ directly into N2,

AOPs are capable of decomposing ammonia-derived nitrogen only when the pH of water


Page 21 of 36


456

and the abiotic process does not significantly depend on operating conditions in terms of

457

treatment performance. In particular, the treatability tests performed in this study confirmed that

458

the high levels of NH3/NH4+ (ca. 4000 mg/L) could rapidly decompose in the EC process. Note

459

that the two-stage biological process is not capable of removing highly concentrated NH3/NH4+

460

since even low concentrations of free ammonia inhibit the biological nitrogen treatment process;

461

the threshold inhibition values of free ammonia on nitrosomonas and nitrobacter (nitrifying

462

bacteria) are known to be ca. 10 mg/L and 0.1 mg/L, respectively (equivalent to 1000 mg/L and

463

20 mg/L of NH4+ plus NH3 at pH 7).58 High salinity in the ADFW allowed the EC process to

464

treat NH3/NH4+ without external supply of NaCl. On the other hand, it is highly probable that the

465

high salt content hampers bacterial growth and metabolism, likely disrupting the biological

466

removal of ammonia.

467

The possible drawback of the sequential E-Fenton/EC process lies in the release of Fe2+

468

and Fe3+ at high concentrations after the operation of anodic Fenton process. To address the

469

concern, the peroxi-coagulation can be considered as a post-treatment step59; soluble iron species

470

are transformed into iron oxides by adjusting the pH of the effluent to a weakly acidic or neutral

471

range, and the produced iron oxides can further contribute to removal of organic and inorganic

472

contaminants through the coagulation mechanism. The EC process as a post-treatment step was

473

demonstrated to directly convert highly-concentrated ammonia into dinitrogen gas in the ADFW,

474

but NO3− level in the effluent corresponded to ca. 12% of the initial NH4+ (ca. 400 mg/L). To

475

tackle nitrate contamination of the effluent, further study is required to develop the cathodes for

476

reductively converting NO3− into N2; the appropriate choice of cathodes (e.g., Fe,60 Cu-Zn61)

477

enables the EC system to simultaneously achieve anodic oxidation of ammonia and cathodic

478

reduction of NO3− without installing an additional treatment step.



Page 22 of 36

479

Acknowledgements

480

This study was supported by a grant from the National Research Foundation of Korea, funded by

481

the Ministry of Science, ICT, and Future Planning (No. 2016M3A7B4909318) and by the Korea

482

Ministry of Environment as the “Global Top Project” (Project No.: 2016002190003).

483

Supporting Information Available.

484

Description of chemicals used in this study (Text S1), SEM images, EDS spectra, and XRD

485

spectra of three commercially-available DSA-type electrodes (Figures S1-S3), Time-dependent

486

changes in the open circuit potentials (Eoc) of anodes and cathodes and potential differences of

487

two plate electrodes in the E-Fenton and EC systems (Figure S4), E-Fenton oxidation of phenol

488

at low feeding rates of Fe2+ and H2O2 (Figure S5), Hydroxylation of benzoic acid and oxidative

489

conversion of methanol to formaldehyde during the E-Fenton process operated at low feeding

490

rates of Fe2+ and H2O2 (Figure S6), Multi-activity assessment of the E-Fenton systems performed

491

at low and high feeding rates of Fe2+ and H2O2 (Figure S7), Effects of electrolyte concentration

492

and type on electrochemical oxidation of NH4+ (Figure S8), Effect of current density on NO3−

493

formation during the electrochemical oxidation of NH4+ (Figure S9), Total dissolved nitrogen

494

measurement during the electrochemical oxidation of NH4+ (Figure S10), NDMA measurement

495

during the electrochemical oxidation of aqueous solutions containing NH4+ alone, NH4+/phenol

496

mixture, and NH4+/phenol/DMA mixture (Figures S11-S12), Mass spectra of gas samples after

497

60 min and 90 min electrolysis of aqueous

498

efficiency of the three DSAs for the electrochemical chlorination of NH4+ at the current density

499

of 150 mA/cm2 (Figure S14), Production of NO3− with increasing current density when the

500

electrochemical chlorination process was used as a pre-treatment step (Figure S15) and as a post-

501

treatment step (Figure S16), Kinetic enhancement in E-Fenton oxidation of benzoic acid, and the

15

NH4+ solutions (Figure S13), Comparison of the


Page 23 of 36


502

generation of active chlorine species when operating the systems with double pairs of anodes and

503

cathodes (Figure S17). This information is available free of charge via the Internet at

504

http://pubs.acs.org/.

505

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aqueous solution. J Phys Chem Ref Data 1988, 17, (2), 513-886.

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44. Zhou, W.; Zhao, H. Q.; Gao, J. H.; Meng, X. X.; Wu, S. H.; Qin, Y. K., Influence of a

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reagents addition strategy on the Fenton oxidation of rhodamine B: control of the competitive

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reaction of •OH. Rsc Adv 2016, 6, (110), 108791-108800.

Sun, L. Z.; Bolton, J. R., Determination of the quantum yield for the photochemical

Wardman, P., Reduction potentials of one-electron couples involving free radicals in


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619

45. Jeong, J.; Kim, C.; Yoon, J., The effect of electrode material on the generation of oxidants

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and microbial inactivation in the electrochemical disinfection processes. Water Res 2009, 43, (4),

621

895-901.

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46. Kapalka, A.; Katsaounis, A.; Michels, N. L.; Leonidova, A.; Souentie, S.; Comninellis, C.;

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Udert, K. M., Ammonia oxidation to nitrogen mediated by electrogenerated active chlorine on

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Ti/PtOx-IrO2. Electrochem Commun 2010, 12, (9), 1203-1205.

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47. Pressley, T. A.; Roan, S. G.; Bishop, D. F., Ammonia-nitrogen removal by breakpoint

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chlorination. Environ Sci Technol 1972, 6, (7), 622-628.

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48. Xiao, S. H.; Qu, J. H.; Zhao, X.; Liu, H. J.; Wan, D. J., Electrochemical process combined

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with UV light irradiation for synergistic degradation of ammonia in chloride-containing

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solutions. Water Res 2009, 43, (5), 1432-1440.

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49. Jafvert, C. T.; Valentine, R. L., Reaction scheme for the chlorination of ammoniacal water.

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Environ Sci Technol 1992, 26, (3), 577-586.

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50. Cho, K.; Hoffmann, M. R., BixTi1-xOz functionalized heterojunction anode with an enhanced

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reactive chlorine generation efficiency in dilute aqueous solutions. Chem Mater 2015, 27, (6),

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51. Hansen, H. A.; Man, I. C.; Studt, F.; Abild-Pedersen, F.; Bligaard, T.; Rossmeisl, J.,

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Electrochemical chlorine evolution at rutile oxide (110) surfaces. Phys Chem Chem Phys 2010,

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storage conditions and pH on chlorine loss in electrolyzed oxidizing (EO) water. J Agr Food

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Chem 2002, 50, (1), 209-212.

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53. Deborde, M.; von Gunten, U., Reactions of chlorine with inorganic and organic compounds

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during water treatment - Kinetics and mechanisms: A critical review. Water Res 2008, 42, (1-2),

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13-51.

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monochloramine by ferrous iron. Environ Sci Technol 2000, 34, (1), 83-90.

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reactions of monochloramine and dichloramine with selected phenols: Examination of humic

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acid models and water contaminants. Environ Sci Technol 2004, 38, (19), 5022-5029.

Vikesland, P. J.; Valentine, R. L., Reaction pathways involved in the reduction of



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electrodes. J Hazard Mater 2009, 165, (1-3), 838-845.

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UV/H2O2. Water Res 2006, 40, (20), 3695-3704.

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electrolytic advanced oxidation processes for the treatment of textile wastewater and sludge

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667 668 669 670 671 672 673 674 675 676 677 678


Page 29 of 36

679


TABLE 1. Quality parameters of anaerobically-digested food wastewater (ADFW) Quality parameter

Measured value

pH

6.9 – 7.1

CODcr (mg/L)

3210 – 3420

TN (mg/L)

4832 – 4728

TP (mg/L)

65.5 – 67.3

NH4+ (mg/L)

3352 – 3790

SS (mg/L)

6260 – 6400

TS (mg/L)

15580

VS (mg/L)

8820 – 8880

Alkalinity (mg/L as CaCO3)

145

Moisture content (%)

98.4

680 681



700

250 (a)

1.0

Phenol 2+ Fe w / H2 O2

600

200

300

150 0.6 100

0.4

200

0.2

100

0.0

2+

Fe H2O2, Fe2+ Conc. (mM)

TOC (mg/L)

400

Phenol Conc. (C/C0 )

0.8 500

Page 30 of 36

w /o H2O2 2+

H2O2 w / Fe

2+

H2O2 w /o Fe TOC

50

0 0

20

40

60

80

100

120

6

5

0.15

4 0.10 3

2 0.05 1

0

4-Hydroxybenzoic Acid Conc. (mM)

Benzoic Acid, Formaldehyde Conc. (mM)

(b)

Benzoic acid 4-Hydroxybenzoic acid Formaldehyde

0.00 0

20

40

60

80

100

120

Reaction Time (min)

682 683 684

FIGURE 1. (a) Degradation of phenol, reduction of total organic carbon, and generation of Fe2+

685

and H2O2 ([phenol]0 = 10 mM; H2O2 input rate = 1.67 mmol/min; current density = 28.5 mA/cm2;

686

[NaCl]0 = 1 M; pHi = 3.0) and (b) hydroxylation of benzoic acid and oxidative transformation of

687

methanol to formaldehyde during the electro-Fenton process ([benzoic acid]0 = 5 mM;

688

[methanol]0 = 1 M; H2O2 input rate = 1.67 mmol/min; current density = 28.5 mA/cm2; [NaCl]0 =

689

1 M; pHi = 3.0).

690


Page 31 of 36


H2 O2 Input Rate (mmol/min)

Pseudo First-Order Rate Constant, k (m hr -1)

0.8

0.84

1.25

1.67

2.09 Current density H2 O2 input rate

0.6

0.4

0.2

0.0 7.1

691

14.2

28.5

42.8 2

Current Density (mA/cm )

692 693

FIGURE 2. Degradation of phenol with increasing current density and H2O2 feeding rate during

694

the electro-Fenton process ([phenol]0 = 10 mM; H2O2 input rate = 1.67 mmol/min; current

695

density = 28.5 mA/cm2; [NaCl]0 = 1 M; pHi = 3.0). Pseudo first-order rate constants were

696

normalized by the specific surface area, as (7.1 m-1).

697 698 699



2

(a)

4000

Page 32 of 36

250 mA/cm 200 mA/cm2

NH4+ Conc. (mg/L)

2

150 mA/cm 2 100 mA/cm 2 50 mA/cm

3000

2000

1000

0 0

20

40

60

80

100

120

140

160

1500 Free and Conbined Chlorine Conc.(mg/L)

(b)

Free chlorine Combined chlorine

1200

900

600

300

0 0

20

40

60

80

100

120

140

160

Reaction Time (min)

700 701 702

FIGURE 3. (a) Electrochemical oxidation of NH4+ under variable current density conditions

703

([NH4+]0 = 4000 mg/L; [NaCl]0 = 1 M; pHi = 5.0) and (b) time-dependent concentration profiles

704

of free and combined chlorines ([NH4+]0 = 4000 mg/L; current density = 200 mA/cm2; [NaCl]0 =

705

1 M; pHi = 5.0).

706 707


Page 33 of 36


100

(a)

12.5 mA/cm2 2

25 mA/cm Degradation Ef ficiency (%)

2

50 mA/cm 2 75 mA/cm

80

60

40

20

Normalized Pseudo First-Order Rate Constant, k (m hr -1 )

0 Nanopac 0.08

Ko-Electrode Samsung DSA

(b)

pH 3 pH 5 pH 7 pH 9

0.06

0.04

0.02

0.00

2

2

2

2

m m A /c m 5 mA /c m mA /c 25 mA /c 50 m 7 12.5

Current Density

708 709 710

FIGURE 4. (a) Comparison of three commercial DSAs in terms of NH4+ oxidation efficiency

711

([NH4+]0 = 4000 mg/L; [NaCl]0 = 1 M; pHi = 5.0) and (b) effect of initial pH on the rate constant

712

of electrochemical NH3/NH4+ oxidation under variable current density conditions ([NH4+]0 =

713

4000 mg/L; [NaCl]0 = 1 M). Pseudo first-order rate constants were normalized by the specific

714

surface area, as (9.6 m-1).

715



Page 34 of 36

2

200 mA/cm2

42.8 mA/cm (TOC) 2

28.5 mA/cm (TOC) 2 14.2 mA/cm (TOC) 2 7.1 mA/cm (TOC) 28.5 mA/cm2 (NH4 +)

800 (a)

2

(b)

4000

3000

2

25 mA/cm

2500

400

2000

200

NH4 + Conc. (mg/L)

3000

NH4 + Conc. (mg/L)

600

TOC (mg/L)

150 mA/cm 2 100 mA/cm 2 50 mA/cm

2000 1500 1000

1000 500

0

0 0

20

40

60

80

100

0

120

0

20

Reaction Time (min)

40

60

80

100

120

140

160

Reaction Time (min) 2

+

2

+

2

+

200 mA/cm (NH4 ) 150 mA/cm (NH4 ) 100 mA/cm (NH4 ) 2

2

42.8 mA/cm 2 28.5 mA/cm 2 14.2 mA/cm 2 7.1 mA/cm

+

50 mA/cm (NH4 ) 2

+

25 mA/cm (NH4 )

(c)

4000

800 (d)

2

200 mA/cm (TOC)

400

400

2000

200

1000

0 0

716

20

40

60

80

100

120

140

0 160

TOC (mg/L)

300 TOC (mg/L)

NH4 + Conc. (mg/L)

600 3000

200

100

0 0

Reaction Time (min)

20

40

60

80

100

120

Reaction Time (min)

717

FIGURE 5. (a) Electro-Fenton oxidation as a pre-treatment process, (b) electrochemical

718

chlorination as a post-treatment process, (c) electrochemical chlorination as a pre-treatment

719

process, and (d) electro-Fenton oxidation as a post-treatment process under variant current

720

density conditions for concomitant removal of phenol and NH4+ ([phenol]0 = 10 mM; [NH4+]0 =

721

4000 mg/L; H2O2 input rate = 1.67 mmol/min; [NaCl]0 = 1 M; pHi (electro-Fenton) = 3.0; pHi

722

(electrochemical chlorination) = 5.0).


Page 35 of 36


E-Fenton process

EC process

1500

TOC (mg/L)

3000

1000 2000

500

1000

0

Ammonia, Nitrate, Nitrite Conc. (mg/L)

4000

(a)

TOC (E-Fenton) + NH4 (E-Fenton) +

NH4 (EC) -

NO3 (EC) -

NO2 (EC) TOC (EC)

0 0

50

100

150

200

250

300

2000

4000

1500

3000

1000

2000

500

1000

0

0 0

723

Ammonia, Nitrate, Nitrite Conc. (mg/L)

TOC (mg/L)

(b)

50

100

150

200

250

300

Reaction Time (min)

724 725

FIGURE 6. Treatment of organic carbon and NH4+-N in the real digested food wastewater by

726

the coupled electro-Fenton/electrochemical chlorination systems consisting of (a) a single pair of

727

anode/cathode and (b) double pairs of anodes/cathodes (current density = 28.5 mA/cm2 (electro-

728

Fenton); current density = 200 mA/cm2 (electrochemical chlorination); H2O2 input rate = 1.67

729

mmol/min; pHi (electro-Fenton) = 3.0; pHi (electrochemical chlorination) = 5.0).



730

Table of Contents Figure:

731


Page 36 of 36

Sequential Combination of Electro-Fenton and Electrochemical

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