In recent years, more and more refractory and toxic organic contaminants are detected in wastewater, surface water and ground water. Many of these organic pollutants can hardly be degraded by conventional water treatments. Advanced oxidation processes (AOPs), which based on in situ generation of strong oxidants, mainly the generation of hydroxyl radicals ( • OH), have been applied to treat various non-biodegradable organic compounds in water. Conventional Fenton technology is a promising AOP for the treatment of organic contaminants in water. In Fenton process, hydroxyl radical, which is the second strongest oxidant (after fluorine), is formed through Fenton's reaction and then degrade organic pollutants. Similar to conventional Fenton process, transition metal ions (Fe 2+ , Co 2+ , Ag + , etc.) can also activate persulfate (PS) and generate sulfate radicals (SO 4

•-). Sulfate radical is a powerful oxidant and can oxidize most of organic pollutants.

This process is named Fenton-like process. There are some disadvantages existed in conventional Fenton and Fenton-like process. For example, a high concentration of Fe 2+ is required and a large amount of iron sludge is generated. In order to solve these problems, electro-Fenton and sulfate radical-based electro-Fenton-like processes are developed. Fe 2+ can be regenerated via cathodic reduction in electro-Fenton and electro-Fenton-like processes. Therefore, the Fe 2+ concentration used in these processes is much lower than that in Fenton and Fenton-like processes. In this paper, electro-Fenton and sulfate radical-based electro-Fenton-like processes were used to degrade artificial sweeteners and azo dyes. The removal efficiency, the oxidation mechanism, degradation pathway and toxicity evolution of target pollutants were investigated.

(1) A detailed discussion on the oxidative degradation of artificial sweetener aspartame (ASP) in acidic aqueous solution containing catalytic amount of Fe 2+ by using electro-Fenton process is reported. In electro-Fenton process, ASP could be completely removed in a 30 min reaction and the removal of ASP followed pseudo-first-order kinetics.

The increase of Fe 2+ concentration and applied current to certain extent could increase the removal efficiency of ASP, while further increasing the Fe 2+ concentration and applied current could lead to the decrease of removal efficiency. Absolute rate constant of IV hydroxylation reaction of ASP was determined as (5.23 ± 0.02) × 10 9 M -1 S -1 . When boron-doped diamond (BDD) electrode was used as anode, ASP could be completely mineralized in a 360 min treatment. Short-chain aliphatic acids such as oxalic, oxamic and maleic acid were identified as aliphatic intermediates in the electro-Fenton process. The bacteria luminescence inhibition showed the toxicity of ASP solution increased at the beginning of electrolysis, and then it declined until lower than the untreated ASP solution at the end of the reaction.

(2) The removal of artificial sweetener saccharin (SAC) in aqueous solution by electro-Fenton processes was performed. Experiments were carried out in an undivided cylindrical glass cell with a carbon-felt cathode and a DSA, Pt or boron-doped diamond (BDD) anode. The removal of SAC by electrochemically generated hydroxyl radicals followed pseudo-first order kinetics with all the anodes. The absolute rate constant of the SAC hydroxylation reaction was found as (1.85 ± 0.01) × 10 9 M -1 s -1 , which was determined by using the competition kinetic method. The comparative study of TOC removal efficiency during electro-Fenton treatment indicated a higher mineralization rate with BDD than Pt anode. Formation of oxalic, formic, and maleic acids were observed during electro-Fenton process. The evolution of toxicity of SAC and/or its reaction by-products based on the V. fischeri bacteria luminescence inhibition was studied.

(3) The removal of artificial sweetener sucralose (SUC) in aqueous solution by electro-Fenton processes was also performed in this study. Hydroxyl radical ( • OH), a highly powerful oxidizing agent, was generated catalytically via the electrochemically assisted Fenton's reaction in the bulk and water oxidation on the surface of anode. The effect of Fe 2+ concentration and applied current on the mineralization of SUC was evaluated. A higher mineralization rate was obtained with BDD than that of Pt anode.

Mineralization current efficiency (MCE) was calculated and better mineralization current efficiency was achieved at relatively short electrolysis time and low applied current. The concentration of the formed carboxylic acids and released inorganic ion was monitored by ion-exlusion chromatography and ion chromatography (IC). The toxicity of SUC and/or its reaction by-products was investigated based on the V. fischeri bacteria luminescence inhibition.

V [START_REF] Heng | Electro-Fenton removal of Orange II in a divided cell: Reaction mechanism, degradation pathway and toxicity evolution[END_REF] The removal of Orange II by a novel electro/α-FeOOH/peroxydisulfate process is reported in this study. In electro/α-FeOOH/peroxydisulfate process, sulfate radicals were generated by activating peroxydisulfate (PDS) with goethite (α-FeOOH). When combined with electrochemical process, the Fe(III) on the surface of α-FeOOH converts to Fe(II) by cathodic reduction. The effect of initial pH on the decolorization of Orange II was investigated. Response surface methodology (RSM) based on Box-Behnken statistical experiment design (BBD) was applied to analyze the experimental variables.

The positive and negative effects on the decolorization of Orange II were determined. The response surface methodology models were derived based on the results of the pseudo-first-order decolorization rate constant and the response surface plots were developed accordingly. The results indicated that the applied current showed a positive effect on the decolorization rate constant of Orange II. The interaction of α-FeOOH dosage and PDS concentration was significant. The ANOVA results confirmed that the proposed models were accurate and reliable for the analysis of the variables of EC/α-FeOOH/PDS process.

( Water quality and availability is a challenging problem facing our society all over the world [START_REF] Rivas | Wastewater recycling: Application of ozone based treatments to secondary effluents[END_REF][START_REF] Bernal-Martí Nez | Synergy of electrochemical and ozonation processes in industrial wastewater treatment[END_REF]. In recent years, more and more organic contaminants from contaminated soil, agricultural runoff, industrial wastewater and hazardous compounds storage leakage are detected in wastewater, surface water and ground water. The presence of these organic compounds in water poses serious threat to public health since most of them are toxic, endocrine disrupting, mutagenic or potentially carcinogenic to humans, animals and aquatic life in general [START_REF] Babuponnusami | A review on Fenton and improvements to the Fenton process for wastewater treatment[END_REF]. Many organic pollutants are considered as toxic and detrimental even when present at very low concentrations. Moreover, many organic contaminants cannot be eliminated by conventional physical separation methods or cannot be degraded by biological processes due to the recalcitrant nature of the pollutants present (Pé rez et al. 2002). Therefore, more powerful wastewater treatment methods are required.

Advance oxidation processes (AOPs) have been proved to be effective for the degradation of many toxic/persistent organic contaminants from aqueous medium such as coloring matters, pesticides, artificial sweeteners and pharmaceuticals and personal care products [START_REF] Kim | Photodegradation characteristics of PPCPs in water with UV treatment[END_REF][START_REF] Wang | Ozonation combined with ultrasound for the degradation of tetracycline in a rectangular air-lift reactor[END_REF][START_REF] Calza | Chemometric assessment and investigation of mechanism involved in photo-Fenton and TiO 2 photocatalytic degradation of the artificial sweetener sucralose in aqueous media[END_REF][START_REF] Janin | Solar photocatalytic mineralization of 2,4-dichlorophenol and mixtures of pesticides: Kinetic model of mineralization[END_REF][START_REF] El-Ghenymy | Decolorization and mineralization of Orange G azo dye solutions by anodic oxidation with a boron-doped diamond anode in divided and undivided tank reactors[END_REF][START_REF] Rodrigo | Electrochemically assisted remediation of pesticides in soils and water: A Review[END_REF][START_REF] Wang | Degradation of Acid Orange 7 in aqueous solution by zero-valent aluminum under ultrasonic irradiation[END_REF]. Moreover, AOPs have been successfully used as pretreatment methods to reduce the concentration of toxic organic compounds that recalcitrant to biological wastewater treatments [START_REF] Stasinakis | Use of selected advanced oxidation processes (AOPs) for wastewater treatment-a mini review[END_REF]. AOPs are based on the in-situ generation of hydroxyl radical ( • OH, (Evgenidou et al. 2007;[START_REF] Salazar | Treatment of a mixture of chloromethoxyphenols in hypochlorite medium by electrochemical AOPs as an alternative for the remediation of pulp and paper mill process waters[END_REF][START_REF] Oturan | Advanced oxidation processes in water/wastewater treatment: Principles and applications. A review[END_REF], which is the second strongest oxidizing agent after fluorine (Mousset et al. 2014a). AOPs are especially efficient for degrading aromatic molecules due to the electrophilic aromatic substitution of hydroxyl radical which then lead to the opening of the aromatic ring (Brillas et al. 2009;Mousset et al. 2014b). The most commonly used AOPs for the removal of organic pollutants from water is based on the Fenton's reagent (an aqueous mixture of Fe 2+ and H 2 O 2 that produces hydroxyl radicals) [START_REF] Babuponnusami | Advanced oxidation of phenol: A comparison between Fenton, electro-Fenton, sono-electro-Fenton and photo-electro-Fenton processes[END_REF][START_REF] Basturk | Advanced oxidation of Reactive Blue 181 solution: A comparison between Fenton and Sono-Fenton Process[END_REF]. It has been used

E 0 ( • OH/H 2 O)= 2.80 V/SHE)
as an attractive and effective technology for the degradation of various organic pollutants (Hermosilla et al. 2009a;Hermosilla et al. 2009b;[START_REF] Zazo | Semicontinuous Fenton oxidation of phenol in aqueous solution. A kinetic study[END_REF]Hermosilla et al. 2012) because of the lack of toxicity of the reagents, eventually leaving no residues and the simplicity of the technology [START_REF] Sun | A kinetic study on the degradation of p-nitroaniline by Fenton oxidation process[END_REF]Nidheesh and Gandhimathi 2012). However, the conventional Fenton process has the disadvantages of requirement of high Fe 2+ concentration and the formation of Fe sludge which limit its application (Nidheesh and Gandhimathi 2012).

Electro-Fenton process, in which H 2 O 2 is produced electrochemically and Fe 2+ can be regenerated throughout the process, overcomes these disadvantages of conventional Fenton process. Compared to the conventional Fenton process, the electro-Fenton process has the advantage of regeneration of Fe 2+ and avoiding the storing and transport of the H 2 O 2 . Moreover, electricity as a clean energy source is used in the process, so the overall process does not create secondary pollutants [START_REF] Jiang | Progress and prospect in electro-Fenton process for wastewater treatment[END_REF]Brillas et al. 2009). The electro-Fenton process has been successfully applied in wastewater treatment and the process efficiency has been experimentally confirmed by different authors for the treatment of media containing pesticides [START_REF] Oturan | Degradation of pesticides in aqueous medium by electro-Fenton and related methods. A review[END_REF][START_REF] Yatmaz | Degradation of pesticide monochrotophos from aqueous solutions by electrochemical methods[END_REF][START_REF] Abdessalem | Treatment of a mixture of three pesticides by photo-and electro-Fenton processes[END_REF], herbicides [START_REF] Kaichouh | Degradation of the herbicide imazapyr by Fenton reactions[END_REF][START_REF] Da Pozzo | Removal of the herbicide amitrole from water by anodic oxidation and electro-Fenton[END_REF]Garcí a et al. 2014), synthetic dyes (Cruz-Gonzá lez et al. 2010;Ghoneim et al. 2011;[START_REF] Kourdali | Degradation of direct yellow 9 by electro-Fenton: Process study and optimization and, monitoring of treated water toxicity using catalase[END_REF], pharmaceticals (El-Ghenymy et al. 2013;[START_REF] Feng | Degradation of anti-inflammatory drug ketoprofen by electro-oxidation: comparison of electro-Fenton and anodic oxidation processes[END_REF][START_REF] Yahya | Oxidative degradation study on antimicrobial agent ciprofloxacin by electro-fenton process: Kinetics and oxidation products[END_REF], and so on.

As a strong oxidant, persulfate (PS) is another popular reagent in the wastewater treatment because it is more stable than H 2 O 2 and it can be transported over a long distance without alteration. Similar to the conventional Fenton process, PS can be activated by transition metals to generate sulfate radicals (SO 4 •-), which have a high standard redox potential (E 0 = 2.6 V/SHE) (Shukla et al. 2010;[START_REF] Deng | Sulfate radical-advanced oxidation process (SR-AOP) for simultaneous removal of refractory organic contaminants and ammonia in landfill leachate[END_REF] and can oxidize organic compounds into small molecules and carbon dioxide (Gayathri et al. 2010). This process is also named Fenton-like process [START_REF] Fernandez | Photobleaching of Orange II within seconds using the oxone/Co 2+ reagent through Fenton-like chemistry[END_REF][START_REF] Fernandez | Bleaching and photobleaching of Orange II within seconds by the oxone/Co 2+ reagent in Fenton-like processes[END_REF].

Like conventional Fenton process, Fe 2+ is hard to be regenerated after conversion to Fe 3+ in sulfate radical-based Fenton-like process, thus high concentration of Fe 2+ is required and a large amount of iron sludge is produced. In order to solve this problem, electrochemical (EC) technology was combined with Fenton-like process (electro-Fenton-like process) and used to treat herbicide 2,4,5-trichlorophenoxyacetic acid [START_REF] Heng | Degradation of clofibric acid in aqueous solution by an EC/Fe 3+ /PMS process[END_REF][START_REF] Heng | Electro-Fenton removal of Orange II in a divided cell: Reaction mechanism, degradation pathway and toxicity evolution[END_REF],5-T) [START_REF] Wang | Degradation of 2,4,5-trichlorophenoxyacetic acid by a novel Electro-Fe(II)/Oxone process using iron sheet as the sacrificial anode[END_REF], azo dye Orange II (Wu et al. 2012), bisphenol A (BPA) (Lin et al. 2013), clofibric acid (CA) (Lin et al. 2014) and landfill leachate [START_REF] Zhang | Removal of COD from landfill leachate by an electro/Fe 2+ /peroxydisulfate process[END_REF]. However, the electro-Fenton-like processes mentioned above are all homogeneous oxidation processes. The removal of metal ions from the effluent requires additional separation and disposal (Zhong et al. 2011). Therefore, various kinds of solid catalysts instead of soluble metal ions have received much attention. Compared with soluble metal ions, solid catalyst can be recovered and reused without being lost along with the effluent. As a result, different kinds of iron metal oxides (such as magnetite and goethite) are employed in the electro-Fenton-like process in this study.

In this work, electro-Fenton and electro-Fenton-like processes in the presence of different oxidants and catalysts were studied. As an emerging contaminant, artificial sweeteners are an essential part of food technology as a group of discrete chemical substances that possess intense sweetness. Synthetic dyes are one of the most common toxic pollutants in the natural environment around the world. These two types of organic compounds were selected as model pollutants in this thesis work.

Research goals and objective

The objective of this work was to study the removal of target pollutants from water using electro-Fenton and electro-Fenton-like processes. The optimal oxidation conditions were identified and the oxidation mechanisms were examined. To meet these goals, the following specific objectives were defined:

1. Conduct experiments to investigate the effect of the operation parameters on the degradation and mineralization of artificial sweetener aspartame (ASP) by electro-Fenton process. The concentration of inorganic ions and carboxylic acid released during the treatment was monitored.

2. Conduct experiments to investigate the degradation and mineralization of artificial sweetener saccharin (SAC) using electro-Fenton process. The comparison among different anode materials on the degradation and mineralization was investigated.

Moreover, the toxicity of SAC and its byproducts during electro-Fenton process was determined.

3. Conduct experiment to investigate the mineralization of artificial sweetener sucralose (SUC) using electro-Fenton process. The influence of the operation parameters on the mineralization of SUC was investigated. The Mineralization current efficiency (MCE) was calculated to evaluate the mineralization capacity.

4. Fe 2+ was replaced by goethite to activate PS and degraded azo dye Orange II in sulfate radical-based electro-Fenton-like process. The response surface methodology models were derived based on the results of the pseudo-first-order decolorization rate constant and the response surface plots were developed accordingly.

5. Magnetite was used as heterogeneous catalyst to degrade Orange II in sulfate radical-based electro-Fenton-like process. The surface properties of Fe 3 O 4 before and after reaction were investigated. The intermediates of Orange II during the treatment were identified and a plausible degradation pathway was proposed. The change of acute toxicity during the treatment was investigated.

Structure of the thesis

This dissertation is composed of eight chapters.

Chapter 1 is an introduction, including background information of electro-Fenton and electro-Fenton-like processes and the chosen organic pollutants, research objectives and organization of this thesis.

Chapter 2 presents a review of recent work related to this study. Four parts are described: Finally, the thesis manuscript ends with Chapter 8 entitled "General conclusions and pespectives".

Photocatalytic oxidation of methyl parathion over TiO 2 and ZnO suspensions.

Catalysis Today 124 [START_REF] Heng | Degradation of bisphenol A in aqueous solution by a novel electro/Fe 3+ /peroxydisulfate process[END_REF][START_REF] Heng | Electro-Fenton removal of Orange II in a divided cell: Reaction mechanism, degradation pathway and toxicity evolution[END_REF], 156-162.

Feng, L., Oturan, N., van Hullebusch, E.D., Esposito, G. and Oturan, M.A. (2014). Hermosilla, D., Merayo, N., Ordóñez, R. and Blanco, Á. (2012). Optimization of conventional Fenton and ultraviolet-assisted oxidation processes for the treatment of reverse osmosis retentate from a paper mill. Waste Management 32 [START_REF] Heng Lin | Treatment of organic pollutants using electro-Fenton and electro-Fenton-like process in aqueous solution[END_REF], 1236-1243.

Degradation

Artificial sweeteners

Artificial sweeteners are one of the most important and interesting classes of emerging contaminants. Artificial sweeteners have been classified as nutritive and non-nutritive depending on whether they are a source of calories [START_REF] Whitehouse | The potential toxicity of artificial sweeteners[END_REF]. The nutritive sweeteners include the monosaccharide polyols (e.g. sorbitol, mannitol, and xylitol) and the disaccharide polyols (e.g. maltitol and lactitol). They are approximately equivalent to sucrose in sweetness [START_REF] Dills | Sugar alcohols as bulk sweeteners[END_REF]. The non-nutritive sweeteners, better known as artificial sweeteners, include substances from several different chemical classes that interact with taste receptors and typically exceed the sweetness of sucrose by a factor of 30 to 13,000 times [START_REF] Whitehouse | The potential toxicity of artificial sweeteners[END_REF]. The most popular artificial sweeteners are aspartame (ASP), saccharin (SAC), sucralose (SUC) and acesulfame. The chemical structure and main characteristics of these commonly used artificial sweeteners were presented in Table 2-1.

Artificial sweeteners are water contaminants that are highly specific to wastewater.

Different from other emerging trace contaminants, such as pharmaceuticals and personal care products (PPCPs), artificial sweeteners have been considered in environmental sciences only recently [START_REF] Loos | EU-wide survey of polar organic persistent pollutants in European river waters[END_REF][START_REF] Mead | Occurrence of the artificial sweetener sucralose in coastal and marine waters of the United States[END_REF][START_REF] Richardson | Environmental mass spectrometry: Emerging contaminants and current issues[END_REF][START_REF] Richardson | Water analysis: Emerging contaminants and current issues[END_REF][START_REF] Lange | Artificial sweeteners-a recently recognized class of emerging environmental contaminants: A review[END_REF]. Excretion after human consumption is one of the major source of artificial sweeteners in the environment [START_REF] Kokotou | Artificial sweeteners as emerging pollutants in the environment: Analytical methodologies and environmental impact[END_REF]. The artificial sweeteners can also enter into wastewater treatment plants from households and industries and they eventually reside in the receiving environmental bodies from effluents [START_REF] Houtman | Emerging contaminants in surface waters and their relevance for the production of drinking water in Europe[END_REF]. Artificial sweeteners are extensively tested for potential adverse health effects on humans because they are used as food additives [START_REF] Schiffman | Sweeteners: State of knowledge review[END_REF][START_REF] Kroger | Low-calorie sweeteners and other sugar substitutes: A review of the safety issues[END_REF][START_REF] En Humanos | Toxicity of sucralose in humans: A review[END_REF]. Although the measured concentrations of some artificial sweeteners range up to microgram per liter levels in surface water, groundwater and drinking water, there is a huge safety margin regarding potential adverse health effects [START_REF] Lange | Artificial sweeteners-a recently recognized class of emerging environmental contaminants: A review[END_REF]. Acceptable daily intake value of artificial sweeteners is 5 mg/kg of body weight per day and are thus three to four orders of magnitude above the maximum possible daily human intake by drinking water. Adverse human health effects for the application of artificial sweeteners have been reported in several studies [START_REF] Lange | Artificial sweeteners-a recently recognized class of emerging environmental contaminants: A review[END_REF]Toth et al. 2012). The long-term health effects resulting from the chronic exposure to low levels of these compounds are largely unknown [START_REF] Mawhinney | Artificial sweetener sucralose in U.S. drinking water systems[END_REF].

Aspartame

Aspartame, the best known of the various artificial sweeteners, is a dipeptide methyl ester ( [START_REF] Sardesai | Natural and synthetic intense sweeteners[END_REF][START_REF] Kokotou | Artificial sweeteners as emerging pollutants in the environment: Analytical methodologies and environmental impact[END_REF]. ASP was found in all of the surface waters at a concentration up to 0.21 mg/L in Tianjin, China (Gan et al. 2013). The risks of ASP ingestion would be in the toxicity of its metabolism products. One of its metabolites is methanol which can cause acidoses and blindness (Conceiç ã o et al. 2005). ASP was rarely degraded by AOPs and it was treated by electro-Fenton process for the first time in this study.

Saccharin

Saccharin (SAC), one of the -first generation‖ artificial sweeteners, was discovered accidentally by Fahlberg and Remsen at Johns Hopkins University in 1879 when they were studying the oxidation of o-toluene-sulfonamides [START_REF] Bruno | Study of thermal degradation of aspartame and its products of conversion in sweetener using isothermal thermogravimetry and HPLC[END_REF]. SAC is about 200-700 times sweeter than sucrose (Assumpç ã o et al. 2008;[START_REF] Lange | Artificial sweeteners-a recently recognized class of emerging environmental contaminants: A review[END_REF].

Nowadays SAC is approved in more than 90 countries and is widely applied in many pharmaceutical and dietary products (Filho et al. 2003). For example, in the European Union (E.U.) SAC is used as an additive in animal feed for piglets, pigs, bovines and calves, and it is also the major degradation by-product of certain sulfonylurea herbicides [START_REF] Buerge | Saccharin and other artificial sweeteners in soils: Estimated inputs from agriculture and households, degradation, and leaching to groundwater[END_REF][START_REF] Kokotou | Artificial sweeteners as emerging pollutants in the environment: Analytical methodologies and environmental impact[END_REF]. SAC was slowly absorbed and not metabolized by the human organism, so it is consequently an appropriate artificial sweetener for diabetics (Filho et al. 2003). SAC has been detected in municipal wastewaters and in rivers in many countries, such as Germany, Switzerland and China (Gan et al. 2013;Van Stempvoort et al. 2011). For example, SAC was detected in surface waters and ranged from 50 ng/L to 0.21 mg/L in Tianjin, China (Gan et al. 2013).

It is crucial to find efficient water treat technologies for the removal of SAC from aqueous media. In this work, electro-Fenton process was employed to degrade SAC in water.

Sucralose

Sucralose (SUC),which is a chlorinated disaccharide derived from sucrose, is one of the most popular artificial sweeteners and has shown increasing trend of consumption (Toth et al. 2012). SUC is about 600 times sweeter than the parent molecule, sucrose (Grotz and Munro 2009;[START_REF] Sharma | Oxidation of artificial sweetener sucralose by advanced oxidation processes: a review[END_REF]. Nowadays, SUC is approved in more than 80 countries and is used widely in food products and pharmaceuticals [START_REF] Sharma | Oxidation of artificial sweetener sucralose by advanced oxidation processes: a review[END_REF]. SUC is not metabolized by the human body due to the orientation of the glycosidic linkage, and consequently, up to 92% of the consumed SUC is excreted unchanged in both urine and feces [START_REF] Sharma | Oxidation of artificial sweetener sucralose by advanced oxidation processes: a review[END_REF]Toth et al. 2012). SUC assesses potential adverse effects on health, since it provokes symptoms, such as, increase in blindness, mineralization of pelvic area and epithelial hyperplasia [START_REF] Calza | Chemometric assessment and investigation of mechanism involved in photo-Fenton and TiO 2 photocatalytic degradation of the artificial sweetener sucralose in aqueous media[END_REF].

Nowadays, SUC has been detected in wastewaters and rivers in many countries, such as America, France, Italy and China [START_REF] Calza | Chemometric assessment and investigation of mechanism involved in photo-Fenton and TiO 2 photocatalytic degradation of the artificial sweetener sucralose in aqueous media[END_REF]Gan et al. 2013). Recently, SUC is considered as an emerging contaminant the by Environmental Protection Agency (EPA) due to its occurrence in environmental waters and persistence (half-life up to several years) [START_REF] Richardson | Environmental mass spectrometry: Emerging contaminants and current issues[END_REF][START_REF] Richardson | Water analysis: Emerging contaminants and current issues[END_REF][START_REF] Calza | Chemometric assessment and investigation of mechanism involved in photo-Fenton and TiO 2 photocatalytic degradation of the artificial sweetener sucralose in aqueous media[END_REF].

Unfortunately, SUC is hard to degrade by conventional wastewater treatments processes [START_REF] Labare | Microbial cometabolism of sucralose, a chlorinated disaccharide, in environmental samples[END_REF][START_REF] Torres | Fate of sucralose during wastewater treatment[END_REF][START_REF] Sharma | Oxidation of artificial sweetener sucralose by advanced oxidation processes: a review[END_REF]. It is crucial to find more powerful water treat technologies for the removal of SUC from polluted effluents.

Different AOPs have been employed to remove SUC from water, including ozonation [START_REF] Scheurer | Performance of conventional multi-barrier drinking water treatment plants for the removal of four artificial sweeteners[END_REF][START_REF] Soh | Fate of sucralose through environmental and water treatment processes and impact on plant indicator species[END_REF], ultraviolet (UV) irradiation [START_REF] Soh | Fate of sucralose through environmental and water treatment processes and impact on plant indicator species[END_REF], UV/H 2 O 2 photooxidation (Keen and Linden 2013), photo-Fenton process [START_REF] Calza | Chemometric assessment and investigation of mechanism involved in photo-Fenton and TiO 2 photocatalytic degradation of the artificial sweetener sucralose in aqueous media[END_REF] and TiO 2 photocatalysis process [START_REF] Calza | Chemometric assessment and investigation of mechanism involved in photo-Fenton and TiO 2 photocatalytic degradation of the artificial sweetener sucralose in aqueous media[END_REF]. When treated by 100 μM ozone, 6% SUC remained after 60 min [START_REF] Soh | Fate of sucralose through environmental and water treatment processes and impact on plant indicator species[END_REF]). Since SUC does not have any evident sites for direct oxidation by ozone, the removal of SUC was mainly caused by hydroxyl radicals generated in the oxidative system [START_REF] Soh | Fate of sucralose through environmental and water treatment processes and impact on plant indicator species[END_REF][START_REF] Sharma | Oxidation of artificial sweetener sucralose by advanced oxidation processes: a review[END_REF]. In a photochemical reactor, SUC did not degrade after 5 h UV exposure [START_REF] Soh | Fate of sucralose through environmental and water treatment processes and impact on plant indicator species[END_REF]) and it also did not decay significantly when UV light irradiation lasted for 24 h [START_REF] Torres | Fate of sucralose during wastewater treatment[END_REF], because SUC has low molar extinction coefficient in the UV region [START_REF] Sharma | Oxidation of artificial sweetener sucralose by advanced oxidation processes: a review[END_REF]. When H 2 O 2 was added to UV irradiation system, the removal of SUC was accelerated and 500 μg/L SUC could be almost totally removed when UV fluence was 4000 mJ/cm 2 (Keen and Linden 2013). When photo-Fenton and TiO 2 photocatalysis processes were introduced to treat SUC, the removal efficiency of 15 mg/L SUC were 97.1% and 88.9%, respectively [START_REF] Calza | Chemometric assessment and investigation of mechanism involved in photo-Fenton and TiO 2 photocatalytic degradation of the artificial sweetener sucralose in aqueous media[END_REF]. SUC was removed from aqueous media by electro-Fenton process for the first time in this study.

Synthetic dyes

Nowadays, there has been an increasing use of synthetic dyes in a large amount of industrial areas such as the textile, leather goods, pharmaceutical industry, food industry and other chemical usages (Forgacs et al. 2004;[START_REF] Hai | Hybrid treatment systems for dye wastewater[END_REF][START_REF] Soon | Heterogeneous catalytic treatment of synthetic dyes in aqueous media using Fenton and photo-assisted Fenton process[END_REF]. In 1856, the world's first commercially synthetic dye was discovered accidentally by William Henry Perkin [START_REF] Saratale | Bacterial decolorization and degradation of azo dyes: A review[END_REF]. Such dyes are defined as colored substances which can give fibers a permanent color. This color is able to resist fading upon exposure to sweat, light, water and many chemicals, including oxidizing agents and microbial attack [START_REF] Rai | Removal of dyes from the effluent of textile and dyestuff manufacturing industry: A review of emerging techniques with reference to biological treatment[END_REF]. More than ten thousand commercially available dyes were developed and used in manufacturing by the end of the 19th century [START_REF] Robinson | Remediation of dyes in textile effluent: a critical review on current treatment technologies with a proposed alternative[END_REF]. The growth of the worldwide textile industry has been accompanied by a rise in pollution due to wastewater contaminated with dyestuff [START_REF] Parikh | Textile dye decolorization using cyanobacteria[END_REF]. The water consumption and the wastewater generation from the textile industry (dry processing mill and woven fabric finishing mills) depend on the processing operations employed during the conversion of fiber to textile fabric [START_REF] Dhanve | Biodegradation of diazo reactive dye Navy blue HE2R (Reactive blue 172) by an isolated Exiguobacterium sp. RD3[END_REF]. The textile industry is one of the greatest generators of liquid effluent pollutants which attributed to the high quantities of water used in the dyeing processes. Moreover, the processing stages and types of synthetic dyes applied during this conversion determine the variable wastewater characteristics in terms of pH, dissolved oxygen, organic and inorganic chemical content [START_REF] Banat | Microbial decolorization of textile-dyecontaining effluents: A review[END_REF]. It is estimated that over 7 × 10 5 tons of textile dyes are discharged in such industrial effluents every year [START_REF] Robinson | Remediation of dyes in textile effluent: a critical review on current treatment technologies with a proposed alternative[END_REF][START_REF] Brillas | Decontamination of wastewaters containing synthetic organic dyes by electrochemical methods. An updated review[END_REF].

Dyes can be classified as Acid, Basic, Direct, Vat, Sulfur, Reactive, Disperse, metal complexes, etc., including antraquinone, indigoide, triphenylmethyl, xanthene and phthalocyanine derivatives (Martí nez-Huitle and Brillas 2009; Brillas and Martí nez-Huitle 2015). Among these dyes, azo dyes, which are characterized by one or two azo-bonds (-N=N-), account for over 70% of all dyestuffs used worldwide [START_REF] Salazar | Finding the best Fe 2+ /Cu 2+ combination for the solar photoelectro-Fenton treatment of simulated wastewater containing the industrial textile dye Disperse Blue 3[END_REF][START_REF] Saratale | Bacterial decolorization and degradation of azo dyes: A review[END_REF][START_REF] Brillas | Decontamination of wastewaters containing synthetic organic dyes by electrochemical methods. An updated review[END_REF], making them the largest group of synthetic colorants and the most common synthetic dyes released into the environment (Chang et al. 2001;[START_REF] Zhao | HPLC and spectrophotometric analysis of biodegradation of azo dyes by Pleurotus ostreatus[END_REF][START_REF] Saratale | Decolorization and biodegradation of textile dye Navy blue HER by Trichosporon beigelii NCIM-3326[END_REF].

Inappropriate disposal of these dyes contaminated waters consistently would result in an even widely contaminated range of environmental matrixes, including surface, ground, drinking water, and soils. Unfortunately, due to the characteristic of biological resistance and chemical stability, azo dyes can hardly be treated using classical wastewater treatments (Ganesh et al. 1994;[START_REF] Razo-Flores | Biodegradation of selected azo dyes under methanogenic conditions[END_REF]. For this reason, many studies have been focused mainly on the removal of azo dyes from waters [START_REF] Shu | Decolorization effects of six azo dyes by O 3 , UV/O 3 and UV/H 2 O 2 processes[END_REF][START_REF] Lucas | Decolorization of the azo dye Reactive Black 5 by Fenton and photo-Fenton oxidation[END_REF]Habibi and Talebian 2007;[START_REF] Migliorini | Anodic oxidation of wastewater containing the Reactive Orange 16 Dye using heavily boron-doped diamond electrodes[END_REF][START_REF] Yu | Acceleration comparison between Fe 2+ /H 2 O 2 and Co 2+ /oxone for decolouration of azo dyes in homogeneous systems[END_REF]). [START_REF] Ince | Combination of activated carbon adsorption with light-enhanced chemical oxidation via hydrogen peroxide[END_REF]. AOPs are considered as the promising processes for the wastewater treatment in 21 st century [START_REF] Munter | Advanced oxidation processes-current status and prospects[END_REF]).

Advanced oxidation processes (AOPs) in dye removal

Fenton process

The Fenton's chemistry started as early as the end of the nineteenth century, when Henry J. Fenton reported that H 2 O 2 could be activated by Fe(II) salts to oxidize tartaric acid [START_REF] Fenton | LXXIII.-Oxidation of tartaric acid in presence of iron[END_REF]. In a century since then, Fenton and related reactions have become of great interest for their relevance to biological chemistry, synthesis and the chemistry of natural waters [START_REF] Pignatello | Advanced oxidation processes for organic contaminant destruction based on the Fenton reaction and related chemistry[END_REF]). In the 1930s, Haber and Weiss showed that the catalytic decomposition of H 2 O 2 by iron salts obeyed to a complex radical and chain mechanism (Haber and Weiss 1934). Since then, interests in Fenton process has been definitely renewed, and Fenton's reagent was applied to oxidize toxic organics originally appeared in the mid-1960s (Brillas et al. 2009).

The generally accepted mechanism of the

Fe 2+ + H 2 O 2 → Fe 3+ + OH  + • OH k 1 = 76 M 1 s 1 (2-1) Fe 3+ + H 2 O 2 → Fe 2+ + HO 2 • + H + k 2 = 0.001-0.01 M 1 s 1 (2-2) RH + • OH → R • + H 2 O (2-3) R • + Fe 3+ → R + + Fe 2+ (2-4) R + + OH  → ROH (2-5)
Nevertheless, a number of competitive reactions can also occur (Eqs (2-6)-(2-10)), which negatively affect the oxidation process [START_REF] Bautista | An overview of the application of Fenton oxidation to industrial wastewaters treatment[END_REF]. The HO 2 • radical formed in Eq. (2-7) is characterized by a lower oxidization power and, therefore, is significantly less reactive towards organic compounds [START_REF] Rush | Pulse radiolytic studies of the reactions of HO 2 /O 2 with Fe (II)/Fe (III) ions. The reactivity of HO 2 /O 2 with ferric ions and its implication on the occurrence of the Haber-Weiss reaction[END_REF][START_REF] Trapido | Fenton treatment efficacy for the purification of different kinds of wastewater[END_REF][START_REF] Oturan | Advanced oxidation processes in water/wastewater treatment: Principles and applications. A review[END_REF].

Fe 2+ + • OH → Fe 3+ + OH  k 3 = 3.20 × 10 8 M 1 s 1 (2-6) H 2 O 2 + • OH → HO 2 • + H 2 O k 4 = 2.70 × 10 7 M 1 s 1 (2-7) Fe 2+ + HO 2 • → Fe 3+ + OH  k 5 = 1.30 × 10 6 M 1 s 1 (2-8) Fe 3+ + HO 2 • → Fe 2+ + O 2 + H + k 6 = 1.20 × 10 6 M 1 s 1 (2-9) • OH + • OH → H 2 O 2 k 7 = 5.20 × 10 9 M 1 s 1 (2-10)
Basically, the Fenton process possesses some important advantages for wastewater treatment, such as easy implementation in existing plants, easy-to-handle and relatively inexpensive chemicals, no need for energy input, and so on [START_REF] Bautista | An overview of the application of Fenton oxidation to industrial wastewaters treatment[END_REF][START_REF] Oturan | Advanced oxidation processes in water/wastewater treatment: Principles and applications. A review[END_REF]. It has been employed to degrade various kinds of organic pollutants in water, for example, dyes [START_REF] Muruganandham | Decolourisation of Reactive Orange 4 by Fenton and photo-Fenton oxidation technology[END_REF][START_REF] Sun | Decolorization of an azo dye Orange G in aqueous solution by Fenton oxidation process: Effect of system parameters and kinetic study[END_REF], landfill leachate [START_REF] Deng | Treatment of landfill leachate by the Fenton process[END_REF][START_REF] Hermosilla | Optimizing the treatment of landfill leachate by conventional Fenton and photo-Fenton processes[END_REF]) and phenol [START_REF] Kavitha | The role of ferrous ion in Fenton and photo-Fenton processes for the degradation of phenol[END_REF][START_REF] Zazo | Chemical pathway and kinetics of phenol oxidation by Fenton's reagent[END_REF].

However, Fenton process has some disadvantages which limit its application.

Firstly, high concentration of Fe 2+ is required and large amount of Fe sludge is formed (Chou et al. 1999). The treatment of the sludge-containing Fe ions at the end of the wastewater treatment is expensive and needs large amount of chemicals and manpower (Nidheesh and Gandhimathi 2012). Secondly, Fenton's reaction is limited by a narrow pH range (pH 2.5-3.5) because iron ions will be precipitated at higher pH values (Nidheesh and Gandhimathi 2012). Thirdly, as a liquid solution, the storing and transport of H 2 O 2 are difficult. In order to solve these problems, Electrochemical advanced oxidation processes (EAOPs) based on Fenton's reaction was developed. externally added to the treated solution to degrade the organic pollutants with • OH from Fenton's reaction (Figure 2-1c). In electrochemical peroxidation process, the coagulation of Fe(OH) 3 precipitate formed depending on the pH and the applied current can be an significant alternative route for the degradation of organic pollutants (Brillas et al. 2009).

Electrochemical

Fe → Fe 2+ + 2e  (2-14)
Electrochemical peroxidation process was employed to treat hexachlorbenzene (HCB) in aqueous solution by Xie et al [START_REF] Xie | Oxidation of hexachlorbenzene by the direct Fenton and EF-Feox method[END_REF]. 96.96% of the initial HCB could be removed in 3 h (Table 2-2). [START_REF] Brillas | Aniline mineralization by AOP's: anodic oxidation, photocatalysis, electro-Fenton and photoelectro-Fenton processes[END_REF][START_REF] Mansour | Relevance of a combined process coupling electro-Fenton and biological treatment for the remediation of sulfamethazine solutions -Application to an industrial pharmaceutical effluent[END_REF][START_REF] Thirugnanasambandham | Modeling of by-product recovery and performance evaluation of Electro-Fenton treatment technique to treat poultry wastewater[END_REF]Zhang et al. 2015a). The major advantage of using undivided cells is that the electrolysis requires lower cell voltage because the voltage penalty of the separator of divided cells is avoided (Brillas et al. 2009). Nidheesh et al. 2014 degraded Rhodamine B (RhB) in a 1000 mL undivided cylindrical glass cell of 10.4 cm diameter containing 750 mL RhB solution (10 mg/L) by using electro-Fenton process. Two graphite plate with the same dimensions (5 × 5 cm) were used as anode and cathode. The results indicated RhB can be totally removed in a 180 min reaction.

However, reactive oxygen species and other weaker oxidants can also be generated at the anode in the undivided cell, which can complicate the electro-Fenton degradation process (Brillas et al. 2009). Therefore, it is harder to discuss the reaction mechanism in an undivided cell. Divided cells, in which cathode and anode are usually separated by 2014 treated pesticide imidacloprid using a heterogeneous electro-Fenton system with ironalginate gel beads (EF-FeAB). The removal efficiency of 100 mg/L imidacloprid was 95% in a 120 min reaction.

Sulfate radical-based Fenton-like process

In recent years, activated persulfate (PS) oxidation is an emerging AOP for the treatment of organic pollutants [START_REF] Rastogi | Effect of inorganic, synthetic and naturally occurring chelating agents on Fe(II) mediated advanced oxidation of chlorophenols[END_REF][START_REF] Tsitonaki | In situ chemical oxidation of contaminated soil and groundwater using persulfate: A Catalysis B[END_REF] , E 0 = 2.01 V/SHE) and peroxymonosulfate (PMS, HSO 5

 , E 0 = 1.82 V/SHE). PS can be activated by transition metals (Fe 2+ , Co 2+ , Ag + , etc.) [START_REF] Romero | Diuron abatement using activated persulphate: Effect of pH, Fe(II) and oxidant dosage[END_REF]Zhao et al. 2010;[START_REF] Pagano | Peroxymonosulfate-Co(II) oxidation system for the removal of the non-ionic surfactant Brij 35 from aqueous solution[END_REF]Wang and Chu 2012;[START_REF] Long | Degradation of toluene by a selective ferrous ion activated persulfate oxidation process[END_REF] to generate sulfate radicals (SO 4 •-), which are promising in the treatment of organic contaminants in water (E 0 = 2.6 V/SHE) (Shukla et al. 2010;Cai et al. 2014;[START_REF] Wang | Degradation of Acid Orange 7 in aqueous solution by zero-valent aluminum under ultrasonic irradiation[END_REF]. Transition metals activated PS process is similar to conventional

Fenton process, so it also be named as Fenton-like process by Fernandez et al [START_REF] Fernandez | Photobleaching of Orange II within seconds using the oxone/Co 2+ reagent through Fenton-like chemistry[END_REF][START_REF] Fernandez | Bleaching and photobleaching of Orange II within seconds by the oxone/Co 2+ reagent in Fenton-like processes[END_REF]). Compared to other transition metals, Fe 2+ is inexpensive, nontoxic and effective, which has been widely used in catalytic oxidation process (Eq. [START_REF] Heng | Degradation of clofibric acid in aqueous solution by an EC/Fe 3+ /PMS process[END_REF][START_REF] Heng | Degradation of bisphenol A in aqueous solution by a novel electro/Fe 3+ /peroxydisulfate process[END_REF][START_REF] Heng | Electro-Fenton removal of Orange II in a divided cell: Reaction mechanism, degradation pathway and toxicity evolution[END_REF][START_REF] Heng | Degradation of Orange II in aqueous solution by a novel electro/Fe 3 O 4 process[END_REF][START_REF] Heng Lin | Treatment of organic pollutants using electro-Fenton and electro-Fenton-like process in aqueous solution[END_REF](7)(8)(9)(10)(11)(12)(13)(14)(15) and(2-16)) (Wang and Chu 2011b;[START_REF] Li | Heterogeneous Fenton oxidation of 2,4-dichlorophenol using iron-based nanoparticles and persulfate system[END_REF]Zhang et al. 2015b). Moreover, due to its unsymmetrical character, PMS might be more easily activated than PDS (Olmez-Hanci et al 2011).

S 2 O 8 2-+ Fe 2+ → Fe 3+ + SO 4 •-+ SO 4 2- (2-15) HSO 5 -+ Fe 2+ → Fe 3+ + SO 4 •-+ OH - (2-16)
Similar to conventional Fenton process, there are some drawbacks encountered in sulfate radical-based Fenton-like process. For example, high ferrous ion dosage is required to activate PS due to the hard regeneration of Fe 2+ . This results in large amount of iron sludge (Vicente et al. 2011). Like EAOPs based Fenton's reaction, electrochemical process was combined with Fenton-like process and solved these problems (electro-Fenton-like process) (Wang and Chu 2011a;Wu et al. 2012).

In sulfate radical-based electro-Fenton-like process, Fe 2+ can be regenerated via cathodic reduction (Eq. (2-12)) of ferric iron ions. In addition, sulfate radicals could be produced via an electron transfer reaction (Eq. [START_REF] Heng | Degradation of clofibric acid in aqueous solution by an EC/Fe 3+ /PMS process[END_REF][START_REF] Heng | Degradation of bisphenol A in aqueous solution by a novel electro/Fe 3+ /peroxydisulfate process[END_REF][START_REF] Heng | Electro-Fenton removal of Orange II in a divided cell: Reaction mechanism, degradation pathway and toxicity evolution[END_REF][START_REF] Heng | Degradation of Orange II in aqueous solution by a novel electro/Fe 3 O 4 process[END_REF][START_REF] Heng Lin | Treatment of organic pollutants using electro-Fenton and electro-Fenton-like process in aqueous solution[END_REF](7)(8)(9)(10)(11)(12)(13)(14)(15)(16)(17) and(2-18)) (Zhao et al. 2010;Zhou et al. 2011). Therefore, the degradation efficiency of organic pollutants was enhanced by coupling EC process and Fe 2+ activated PS process.

S 2 O 8 2-+ e -→ SO 4 •-+ SO 4 2- (2-17) HSO 5 -+ e -→ SO 4 •-+ OH - (2-18)
According to different producing way of the PS and Fe 2+ , electro-Fenton-like process can be divided into three categories.

In the first one, an iron electrode was used as anode and provided soluble ferrous ions continuously through anodic oxidation (Eq. [START_REF] Heng | Degradation of clofibric acid in aqueous solution by an EC/Fe 3+ /PMS process[END_REF][START_REF] Heng | Degradation of bisphenol A in aqueous solution by a novel electro/Fe 3+ /peroxydisulfate process[END_REF][START_REF] Heng | Electro-Fenton removal of Orange II in a divided cell: Reaction mechanism, degradation pathway and toxicity evolution[END_REF][START_REF] Heng | Degradation of Orange II in aqueous solution by a novel electro/Fe 3 O 4 process[END_REF][START_REF] Heng Lin | Treatment of organic pollutants using electro-Fenton and electro-Fenton-like process in aqueous solution[END_REF](7)(8)(9)(10)(11)(12)(13)(14)). Once PS was added into the system, sulfate radicals were generated via the catalysis of ferrous ions (Eq. [START_REF] Heng | Degradation of clofibric acid in aqueous solution by an EC/Fe 3+ /PMS process[END_REF][START_REF] Heng | Degradation of bisphenol A in aqueous solution by a novel electro/Fe 3+ /peroxydisulfate process[END_REF][START_REF] Heng | Electro-Fenton removal of Orange II in a divided cell: Reaction mechanism, degradation pathway and toxicity evolution[END_REF][START_REF] Heng | Degradation of Orange II in aqueous solution by a novel electro/Fe 3 O 4 process[END_REF][START_REF] Heng Lin | Treatment of organic pollutants using electro-Fenton and electro-Fenton-like process in aqueous solution[END_REF](7)(8)(9)(10)(11)(12)(13)(14)(15)(16)(17) or

(2-18)) (Figure 2-3a).
Wang and Chu 2011a proposed an -electro-Fe(II)/Oxone (EFO)‖ process to degrade 2,4,4, Cast gray iron was used as anode and provided Fe 2+ , which could activate persulfate to generate sulfate radicals. 99% of 0.4 mM TCE could be transformed in a 20 min reaction time.

In the second category, both PS and Fe 2+ are added externally to the reactor. Fe 3+ which formed in Eq. ( 2 -Electro/Fe 2+ /peroxydisulfate‖ process was also applied to the treatment of landfill leachate by Zhang et al [START_REF] Zhang | Removal of COD from landfill leachate by an electro/Fe 2+ /peroxydisulfate process[END_REF]. When PS concentration was 62.5 mM, Fe 2+ concentration was 15.6 mM, j was 13.89 mA/cm 2 and initial pH 3.0, the COD removed by oxidation and coagulation was about 40% and 22%, respectively.

In the third category, Fe 2+ is replaced by Fe 3+ and added to the treated solution because Fe 2+ has the disadvantage of easily oxidized to Fe 3+ by air and ferrous solution should be stored under an acidic condition. Fe 3+ was first reduced to Fe 2+ (Eq. (2-12)) at the cathode and then activated PS to generate sulfate radicals. The regeneration of ferrous ions was enhanced by cathodic reduction (Eq. (2-12)) simultaneously (Figure 2-3c). -Electro/Fe 3+ /PDS‖ and -EC/Fe 3+ /PMS‖ processes were used to treat organic pollutants in water by our team (Lin et al. 2013(Lin et al. , 2014)). -Electro/Fe 3+ /PDS‖ was used to degrade bisphenol A (BPA) which is a known endocrine disrupter and industrial chemical [START_REF] Staples | A review of the environmental fate, effects, and exposures of bisphenol A[END_REF]Cleveland et al. 2014;[START_REF] Olmez-Hanci | S 2 O 8 2-/UV-C and H 2 O 2 /UV-C treatment of Bisphenol A: Assessment of toxicity, estrogenic activity, degradation products and results in real water[END_REF]. BPA could be nearly completely removed in a 60 min treatment time with initial BPA concentration of 0.22 mM, Fe 3+ concentration of 4 mM, PDS concentration of 20 mM, Na 2 SO 4 concentration of 50 mM, initial pH value of 3.0 and the current density of 33.6

mA/cm 2 . The mineralization efficiency reached 94.3% in a 120 min reaction at the same condition (Lin et al. 2013). Clofibric acid (CA), which is the active metabolite of several widely used blood lipid regulators, was treated by -EC/Fe 3+ /PMS‖ process (Lin et al. 2014). When initial CA concentration was 50 mg/L, PMS concentration fixed at 20 mM, Fe 3+ concentration was 2.00 mM, Na 2 SO 4 concentration was 50 mM, initial pH was 4.0 and the current density was 33.6 mA/cm 2 , CA was completely removed after 60 min reaction. All the electro-Fenton-like processes mentioned above are homogeneous processes. They have the same drawbacks as homogeneous electro-Fenton processes. In this study, iron metal oxides were used to replace Fe 2+ and activated PS.

Iron oxides applied in AOPs based on Fenton and Fenton-like process

Iron oxides are abundantly available minerals, presenting in the natural aqueous environment as suspended solid particles, and they also suspend in aerosol, clouds and fogs as fine particles [START_REF] Hou | Shape-controlled nanostructured magnetite-type materials as highly efficient Fenton catalysts[END_REF][START_REF] Rahim Pouran | Review on the application of modified iron oxides as heterogeneous catalysts in Fenton reactions[END_REF]. Recently, iron oxides are widely used in heterogeneous catalysis processes and constitute an attractive alternatives for the decontamination of soils, underground waters, sediments, and industrial effluents because they are natural, abundant, inexpensive, and environmentally friendly [START_REF] Aredes | The removal of arsenic from water using natural iron oxide minerals[END_REF][START_REF] Xu | Use of iron oxide nanomaterials in wastewater treatment: A review[END_REF][START_REF] Rahim Pouran | Review on the application of modified iron oxides as heterogeneous catalysts in Fenton reactions[END_REF].

Nowadays, several iron oxides and modified iron oxides are used to replace Fe 2+ in AOPs based Fenton and Fenton-like process [START_REF] He | Photoreaction of aromatic compounds at α-FeOOH/H 2 O interface in the presence of H 2 O 2 : evidence for organic-goethite surface complex formation[END_REF]Guo et al. 2010;[START_REF] Ji | Efficient performance of porous Fe 2 O 3 in heterogeneous activation of peroxymonosulfate for decolorization of Rhodamine B[END_REF][START_REF] He | Heterogeneous Fenton oxidation of catechol and 4-chlorocatechol catalyzed by nano-Fe 3 O 4 : Role of the interface[END_REF][START_REF] Sun | Enhanced heterogeneous and homogeneous Fenton-like degradation of carbamazepine by nano-Fe 3 O 4 /H 2 O 2 with nitrilotriacetic acid[END_REF]).

Goethite (α-FeOOH)

Goethite, an iron oxyhydroxide mineral with chemical formula of α-Fe III O(OH), is a commonly used iron oxide among Fe(III) bearing minerals for heterogeneous Fenton and Fenton-like based AOPs due to its ability to operate in a wide range of pH, higher stability thermodynamically and being relatively low priced and environmentally friendly (Ortiz de la [START_REF] Ortiz De La Plata | Optical properties of goethite catalyst for heterogeneous photo-Fenton reactions: Comparison with a titanium dioxide catalyst[END_REF]Wang et al. 2010b;[START_REF] Rahim Pouran | Review on the application of modified iron oxides as heterogeneous catalysts in Fenton reactions[END_REF][START_REF] Sable | FeOOH and derived phases: Efficient heterogeneous catalysts for clofibric acid degradation by advanced oxidation processes (AOPs)[END_REF]. Goethite is also used in electro-Fenton process as a substitute of Fe 2+

to react with H 2 O 2 and form • OH through Fenton's reaction, avoiding the additional water pollution caused by the homogeneous catalyst. [START_REF] Expósito | Mineral iron oxides as iron source in electro-Fenton and photoelectro-Fenton mineralization processes[END_REF] applied goethite as active heterogeneous catalyst undergo Fenton's reaction with electrogenerated H 2 O 2 in electro-Fenton process using aniline as the model pollutant.

When initial aniline concentration was 100 mg/L, α-FeOOH dosage fixed at 1 g/L, Na 2 SO 4 concentration was kept at 20 mM, initial pH was 3.0 and the applied current was 250 mA, the TOC removal efficiency of aniline was about 90% in a 25 h reaction.

Magnetite (Fe 3 O 4 )

Magnetite is a spinel iron oxide with chemical formula of (Fe 3+ ) tet [Fe 2+ Fe 3+ ] oct O 4

where Fe 3+ cations occupy equally both octahedral and tetrahedral sites and Fe 2+ cations are placed only in octahedral sites [START_REF] Rahim Pouran | Review on the application of modified iron oxides as heterogeneous catalysts in Fenton reactions[END_REF]. Magnetite has gained considerable attention than other iron oxides in AOPs based Fenton and Fenton-like process due to its unique characteristics: (i) The Fe(II) in its structure may play an significant role as an electron donor to initiate the Fenton's reaction [START_REF] Kwan | Rates of hydroxyl radical generation and organic compound oxidation in mineral-catalyzed Fenton-like systems[END_REF][START_REF] Moura | Efficient use of Fe metal as an electron transfer agent in a heterogeneous Fenton system based on Fe 0 /Fe 3 O 4 composites[END_REF][START_REF] Hou | Shape-controlled nanostructured magnetite-type materials as highly efficient Fenton catalysts[END_REF]); (ii) The octahedral site in the magnetite structure can easily accommodate both Fe(II) and Fe(III), which means that Fe(II) can be reversibly oxidized and reduced in the same structure [START_REF] Moura | Efficient use of Fe metal as an electron transfer agent in a heterogeneous Fenton system based on Fe 0 /Fe 3 O 4 composites[END_REF]); (iii) The production of more active systems by modification in the physico-chemical properties of the magnetite through isostructural substitution of iron by different transition metals [START_REF] Moura | Efficient use of Fe metal as an electron transfer agent in a heterogeneous Fenton system based on Fe 0 /Fe 3 O 4 composites[END_REF][START_REF] Hou | Shape-controlled nanostructured magnetite-type materials as highly efficient Fenton catalysts[END_REF]); (iv) The magnetically easy separation of magnetite catalysts from the reaction system as a result of its magnetic property [START_REF] Ai | Removal of methylene blue from aqueous solution with magnetite loaded multi-wall carbon nanotube: Kinetic, isotherm and mechanism analysis[END_REF]Chun et al. 2012) and (v) Higher dissolution rate of magnetite compared to other iron oxides which lead to higher electron mobility in its spinel structure [START_REF] Litter | Photodissolution of iron oxides. IV. A comparative study on the photodissolution of hematite, magnetite, and maghemite in EDTA media[END_REF][START_REF] Matta | Oxidation of 2,4,6-trinitrotoluene in the presence of different iron-bearing minerals at neutral pH[END_REF].

Chen Separation and Purification Technology 80(3), 626-634.

Introduction

In this study, a detailed discussion on the oxidative degradation of artificial sweetener aspartame (ASP) in acidic aqueous solution containing catalytic amount of Fe 2+ by using electro-Fenton process was reported. The kinetics of ASP degradation by

• OH generated during electro-Fenton process has been investigated. The absolute rate constant of the reaction between ASP and • OH was determined by the competition kinetic method using benzoic acid as a reference competitor (Özcan et al. 2009a, Özcan et al. 2013). The effect of the applied current and Fe 2+ concentration on the degradation and mineralization of ASP was examined. The aliphatic short-chain carboxylic acids released during electro-Fenton process were monitored. The concentration of inorganic ions released into the solution was investigated. The variation of toxicity of ASP solution and its intermediates was assesed by Microtox method. and Fluka. All solutions were prepared with ultrapure water obtained from a Millipore Mill-Q system with resistivity > 18 MΩ cm at room temperature.

Materials and methods

Procedures and equipment

Bulk experiments were carried out at room temperature in a 250 mL undivided cylindrical glass cell of 6 cm diameter containing 220 mL ASP solution. The cathode (working) electrode was a 87.5 cm 2 piece of carbon felt (17.5 cm × 4.5 cm, Carbon-Lorraine, France) which placed on the inner wall of the cell covering the totality of the internal perimeter. The anode (counter) electrode was a cylindrical Pt mesh (4.5 cm height, i.d. = 3.1 cm, Platecxis, France) or a 24 cm 2 thin-film BDD electrode (CONDIAS GmbH, Germany). The anode was centered in the cell, surrounded by the carbon felt. Electrolysis was conducted under constant current conditions using a Hameg HM8040-3 triple power supply (Germany).

Compressed air was bubbled starting 5 min before electrolysis at about 0.5 L/min through the aqueous solutions to saturate aqueous solution before starting electrolysis and it is maintained during electrolysis to ensure the continuous saturation of oxygen. and set at 3.0 (±0.1), adjusting by the addition of 1 M sulfuric acid.

High performance liquid chromatography (HPLC) analysis

The initial and residual ASP concentration was monitored by HPLC, which consists of a Merck Lachrom liquid chromatograph, equipped with a L-2130 pump and was fitted with a Purospher RP-18, 5 μm, 25 cm × 4.6 mm (i.d.) column at 40 C, and coupled with a L-2400 UV detector at maximum absorption wavelength of 215 nm for ASP. The analysis was performed using a 70:30 (v/v) water (0.1% acetic acid)/methanol (0.1% acetic acid) isocratic solvent mixture as mobile phase at a flow rate of 0.5 mL/min. The injection volume was 20 μL.

The short-chain carboxylic acids were identified and quantified by ion-exclusion HPLC consisting of a Merck Lachrom liquid chromatograph equipped with a L-7100 pump, a Supelcogel H column (250 mm × 4.6 mm, 9 μm (i.d.) particle size) and a L-7455 photodiode array detector at a selected wavelength of 220 nm. 0.1% H 2 SO 4 solution was used as mobile phase with a flow rate of 0.2 mL/min. Calibration curves were achieved using standard solutions of related carboxylic acids. The identification of the carboxylic acids was performed by the comparison between the retention time (t R )

and internal standard addition method using standard solutions.

Ion chromatography (IC) analysis

Inorganic

Total organic carbon (TOC) analysis

The TOC of the initial and electrolyzed samples were determined by Shimatzu TOC-V CSH analyser according to the combustion catalytic oxidation method at 680 C.

The carrier gas was oxygen with a flow rate of 150 mL/min. Platinum was used as catalyst to carry out the combustion at 650 °C instead of 900 °C. The injection volume was 50 μL. Calibration of the analyser was obtained with potassium hydrogen phthalate standards.

Toxicity measurements

The evolution of the toxicity of treated solutions during the treatment was assessed by Microtox method according to the international standard process (OIN 11348-3) using a Microtox ® model 500. This measurements performed with the bio-luminescence marine bacteria V. fischeri [START_REF] Buerge | Ubiquitous occurrence of the artificial sweetener acesulfame in the aquatic environment: An ideal chemical marker of domestic wastewater in groundwater[END_REF][START_REF] Buerge | Saccharin and other artificial sweeteners in soils: Estimated inputs from agriculture and households, degradation, and leaching to groundwater[END_REF][START_REF] Bokare | Review of iron-free Fenton-like systems for activating H 2 O 2 in advanced oxidation processes[END_REF], provided by Hach Lange France SAS. Two values of the inhibition of the luminescence (%) were measured after 5 min and 15 min of exposure to samples at 15 C.

Results and discussion

Effect of catalyst concentration on ASP degradation

A set of electrolysis was carried out with 0.2 mM (58.9 mg/L) ASP in aqueous solutions to determine the influence of the main operating parameters on the ASP degradation in electro-Fenton process. The results obtained show that the oxidative degradation of ASP by electro-Fenton process fits well with the first-order kinetic model and the apparent rate constant values (k app ) calculated accordingly under different operating conditions were given in table 3-1. The effect of catalyst (Fe 2+ ) concentration in ASP oxidation by electro-Fenton process was performed at current-controlled conditions and room temperature with Pt and BDD anode, respectively (Fig. 3-1). The Fe 2+ concentration was varied in the range of 0.050.5 mM in 50 mM Na 2 SO 4 solution, at applied current 200 mA and initial pH 3.0. As can be seen in Fig. 3 With both Pt and BDD anode, the removal efficiency and degradation rate of ASP increased significantly when Fe 2+ concentration increased from 0.05 to 0.2 mM (Fig.

3-1 and Table 3-1). At higher Fe 2+ concentration, a considerable amount of hydroxyl radicals would be produced via Fenton's reaction. These hydroxyl radicals react with ASP immediately, resulting in the increase of ASP degradation rate. However, further increasing the Fe 2+ concentration to 0.5 mM, ASP removal efficiency and rate constant decreased evidently (Fig. 3-1 and Table 3-1). The negative influence of higher Fe 2+

Fe 2+ + • OH → Fe 3+ + OH  (3-2)
Therefore, the reaction (3-2) becomes competitive for OH consumption at higher Fe 2+

and consequently harms the oxidation of ASP.

Effect of applied current on ASP degradation

In electro-Fenton process, the applied current is a crucial parameter for the operational cost and process efficiency (Özcan et al. 2009a), because the formation rate of H 2 O 2 (Eq. [START_REF] Heng | Degradation of bisphenol A in aqueous solution by a novel electro/Fe 3+ /peroxydisulfate process[END_REF][START_REF] Heng | Electro-Fenton removal of Orange II in a divided cell: Reaction mechanism, degradation pathway and toxicity evolution[END_REF]), the regeneration rate of Fe 2+ (Eq. ) and consequently the generation rate of  OH through Fenton's reaction (Eq. (2-1)) are governed by this parameter. The applied current governs also the formation rate (Eq. (3-1)) and the amount of heterogeneous hydroxyl radicals BDD(  OH) or Pt(  OH). Therefore the effect of applied current for the treatment of 0.2 mM ASP in electro-Fenton processes using Pt/carbon-felt (Fig. 3-2a) and BDD/carbon-felt (Fig. 3-2b) cells was examined by using the applied current of 50, 100, 200, 300 and 500 mA.

As depicted in Fig. 3-2a, the total disappearance of ASP became faster at a higher current in Pt/carbon-felt cell. The k app increased from 0.14 to 0.34 min -1 when the applied current increased from 50 to 300 mA (Table 3-1). A higher current can promote both the Fe 2+ regeneration (Eq. ) and the production of H 2 O 2 (Eq. (3-4)) (Lin et al. 2014). However, further increasing the current to 500 mA resulted in the decrease of k app to 0.17 min -1 . The decrease in ASP oxidation efficiency at higher current can be related to the increase of parasitic reactions, such as H 2 evolution reaction (Eq.(3-5)) (Dirany et al. 2012).

Fe 3+ + e  → Fe 2+ (3-3) O 2 + 2H + + 2e  → H 2 O 2 (3-4) 2H 2 O + 2e -→ H 2 + 2OH - (3-5)
When it comes to the BDD/carbon-felt cell, the optimal current for ASP degradation was 200 mA (Fig. 3-2b). Moreover, it can be seen from Fig. 2 that the comparatively better performance of BDD was more pronounced at lower current intensities. When applied current is higher than 200 mA, the positive role of BDD favoring generation of • OH was insignificant (Table 3-1 occurring on carbon-felt cathode (Özcan et al. 2009b). Moreover, the O 2 evolution reaction was more dominate than • OH on the BDD anode at high applied current (Özcan et al. 2009b). (iii) The formation of weak oxidant species like HO

2012)

.

O 2 + 4H + + 4e -→ H 2 O (3-6) H 2 O 2 + • OH → HO 2 • + H 2 O (3-7) HO 2 • + • OH → H 2 O + O 2 (3-8) (a) (b)

Determination of absolute constants for oxidation of ASP by hydroxyl radicals

The absolute rate constant of ASP treatment by hydroxyl radicals during electro-Fenton process was determined by competition kinetics method. Benzoic acid (BA) was selected as standard competitor because the absolute rate constant (k abs,BA ) of the reaction between BA and hydroxyl radicals is well known (k abs,BA = 4.30 × 10 9 M -1 s -1 ) (Özcan et al. 2008;Siré s et al. 2007c;Oturan et al. 2010). Experiments were performed using Pt anode with 0.1 mM ASP solution, Fe 2+ concentration 0.2 mM, applied current 50 mA and initial pH 3.0. The absolute rate constant of the oxidation of ASP was then calculated according to the Eq. [START_REF] Heng | Degradation of bisphenol A in aqueous solution by a novel electro/Fe 3+ /peroxydisulfate process[END_REF][START_REF] Heng | Electro-Fenton removal of Orange II in a divided cell: Reaction mechanism, degradation pathway and toxicity evolution[END_REF][START_REF] Heng | Degradation of Orange II in aqueous solution by a novel electro/Fe 3 O 4 process[END_REF][START_REF] Heng Lin | Treatment of organic pollutants using electro-Fenton and electro-Fenton-like process in aqueous solution[END_REF](7)(8)(9).

Ln ( [ASP] 0 [ASP] t ) = ( k abs,ASP k abs,BA ) Ln ( [BA] 0 [BA] t ) (3-9)
Based on Fig. 3-3, the absolute rate constant for the oxidation reaction of ASP and • OH (k abs,ASP ) was then determined as (5.23 ± 0.02) × 10 9 M -1 s -1 . This value is closed to that reported by Toth et al. 2012 (6.06 ± 0.05) × 10 9 M -1 s -1 ) who determined the absolute rate constant by direct observation of the formation of the cyclohexadienyl radical adduct. For both Pt/carbon-felt and BDD/carbon-felt cell, the TOC removal efficiency decreased as Fe 2+ concentration increased from 0.1 to 0.5 mM. With the increasing Fe 2+ concentration, the percentage of scavenged hydroxyl radicals from Fe 2+ ions increased, so the side reactions between Fe 2+ ions and • OH (Eq. (3-2)) became more significant (Özcan et al. 2008). This reaction becomes particularly important at longer treatment time because of lower concentration of organic matter in the solution. Furthermore, as can be seen from Fig. 34, the TOC removal rate was fast during the first 120 min for all the Fe 2+ concentration in both cells. After 120 min, the TOC removal values gradually decreased and reached a steady state. This is due to the quick decomposition of ASP and its aromatic derivatives which are more easily oxidizable than the ring opening products such as aliphatic acids obtained at longer treatment times (Hammami et al. 2008). The slower mineralization rate observed at longer electrolysis times for both Pt and BDD anodes can also be attributed to the formation of stable ferro-and/or ferri-complexes with carboxylic acids [START_REF] Oturan | Complete destruction of p-Nitrophenol in aqueous medium by electro-Fenton method[END_REF][START_REF] Brillas | Mineralization of herbicide 3,6-dichloro-2-methoxybenzoic acid in aqueous medium by anodic oxidation, electro-Fenton and photoelectro-Fenton[END_REF][START_REF] Guivarch | Degradation of azo dyes in water by Electro-Fenton process[END_REF]Hammami et al. 2008;[START_REF] Haidar | Electrochemical degradation of the antibiotic sulfachloropyridazine by hydroxyl radicals generated at a BDD anode[END_REF].

Even so, it is evident that the use of BDD anode lead to the acceleration of ASP mineralization (Fig. 34). For example, at the end of the 360 min electrolysis, 84.5% and 98.0% of the initial TOC value was removed, in the case of Fe 2+ concentration of The effect of applied current on the mineralization of ASP solution in terms of TOC abatement under the same conditions of Fig. 3-2 was investigated and the results

were shown in Fig. 345. With the raising applied current from 100 to 200 mA, the TOC removal efficiency was found to increase from 72.2% to 81.7% in Pt/carbon-felt cell.

The TOC removal efficiency increased from 93.1% to 97.5% in BDD/carbon-felt cell at the end of 360 min electrolysis under the same conditions. Higher applied current lead to the higher accumulation of hydroxyl radicals in the bulk, and then a greater mineralization efficiency can be achieved because of the simultaneous degradation of ASP and its byproducts (Dirany et al. 2012). Further increase in the applied current to 500 mA, the TOC removal efficiency declined to 71.6% and 93.4% for Pt and BDD anode, respectively. In the treatment of ASP, the application of current higher than 200 mA would increase the extent of parasitic reactions (Eq. ( 3

Identification and evolution of carboxylic acids

It is well known that the cleavage of the benzenic ring of aromatics can lead to the release of short-chain carboxylic acids in electro-Fenton process (Siré s et al. 2007a;Özcan et al. 2009b;[START_REF] Vez | Mineralization of the drug β-blocker atenolol by electro-Fenton and photoelectro-Fenton using an air-diffusion cathode for H 2 O 2 electrogeneration combined with a carbon-felt cathode for Fe 2+ regeneration[END_REF]Dirany et al. 2012;El-Ghenymy et al. 2013). The concentration of carboxylic acids were monitored by ion-exclusion HPLC when 0.2 mM ASP solutions were electrolyzed under the conditions of Figs. 3-4 and 3-5. In this study, oxalic, oxamic and maleic acid at retention time (t R ) of 8.90, 14.26 and 11.33 were observed, respectively, in both Pt/carbon-felt and BDD/carbon-felt cell.

Among these carboxylic acids, the concentration of maleic acid was detected in trace and only exist in the first 10 min in both cells, so changes of maleic acid concentration were not shown in Fig. 3456.

As shown in Fig. 3-6, all the carboxylic acids were generated at the beginning of the electrolysis and followed the accumulation-destruction cycles. A similar behavior has been observed for the treatment of other organic products (Özcan et al. 2009b;Dirany et al. 2012). Oxalic and oxamic acids were remained even at the end of the electrolysis indicating their lower reactivity with hydroxyl radicals and need much more time to destruction (Dirany et al. 2012). It also depicted by comparing Fig. 3-6a and Fig. 3-6b that all the acids were nearly disappeared at the end of electrolysis (360 min) in BDD/carbon-felt cell, in agreement with the mineralization results shown in Fig. 3-4b and Fig. 3-5b. However, oxalic and oxamic acid were still existed in Pt/carbon-felt cell in 360 min reaction, which was corresponding to the residual TOC remaining at the end of mineralization treatments (Fig. 3-4a and Fig. 3-5a).

Materials and methods

Chemicals

Saccharin (1,1-dioxo-1,2-benzothiazol-3-one (or o-benzoic sulfimide), C 7 H 5 CNO 3 S) was purchased from Sigma-Alorich. Analytical grade anhydrous sodium sulfate and ferrous sulfate heptahydrated were obtained from Sigma-Aldrich and Acros Organics, respectively. Regent grade benzoic acid was purchased from Problabo (France) used as the competition substrate in several kinetic experiments. Analytical grade carboxylic acids and other chemicals used for chromatographic analysis were purchased from Acros, Merck, Sigma, Riedel-de Haë n and Fluka. Ultrapure water used for the preparation of the working solutions and HPLC eluting solutions was obtained from a Millipore Milli-Q (simplicity 185) system with resistivity > 18 MΩ cm at room temperature.

Electrochemical apparatus and procedures

Electrolyses were performed at constant current and room temperature using a Hameg HM8040-3 triple power supply (Germany) in an open, cylindrical undivided glass cell of 6 cm diameter and 250 mL capacity containing 220 mL SAC solution.

Electro-Fenton oxidation was conducted using three anodes: a cylindrical Pt mesh (4.5 cm height, i.d. = 3.1 cm, Platecxis, France), a 25 cm 2 thin-film BDD (CONDIAS GmbH, Germany) and a commercial DSA (mixed metal oxide Ti/RuO 2 -IrO 2 , Baoji Xinyu GuangJiDian Limited Liability Company, China). A 87.5 cm 2 piece of carbon felt (17.5 × 5 cm, Carbon-Lorraine, France) was used as cathode.

In all cases, the anode was centered in the electrolytic cell and was surrounded by the cathode that covered the inner wall of the cell. H 2 O 2 was produced from reduction of O 2 dissolved in the solution. Continuous saturation of O 2 at atmospheric pressure was ensured by bubbling compressed air passing through a frit at about 0.5 L/min, starting 5 min before the beginning of the electrolysis. Prior to the electrolysis, a catalytic quantity of ferrous ion and 50 mM Na 2 SO 4 (supporting electrolyte) were added to the SAC solutions with constant stirring by using a magnetic stirrer (IKA, Germany).

The pH of initial solutions (pH 0 ) was set at 3.0 by the addition of 1 M sulfuric acid because this value was reported as the optimal pH value for the electro-Fenton processes (Özcan et al. 2008a). The pH of SAC solutions was measured with a CyberScan pH 1500 pH-meter (Eutech Instrument, USA).

Analytical methods and procedures

The concentration of residual SAC was monitored by high performance liquid chromatography (HPLC), which consist of a Merck Lachrom liquid chromatograph, equipped with a L-2130 pump and fitted with a Purospher RP-18, 5 μm, 25 × 4.6 mm (i.d.) column at 40 C, and coupled with a L-2400 UV detector at maximum absorption wavelength of 218 nm for SAC. The analyses were performed using a phosphoric acid (pH 3.0)/methanol (80:20, v/v) as mobile phase in isocratic mode at a flow rate of 1.0 mL/min. The injection volume was 20 μL.

The short-chain carboxylic acids were identified and quantified by ion-exclusion HPLC using a L-7100 pump, a Merck Lachrom liquid chromatograph equipped with a Supelcogel H column (250 × 4.6 mm, 9 μm) and a L-7455 photodiode array detector at the wavelength of 220 nm. The mobile phase was 0.1% H 2 SO 4 solution and flow rate was fixed to 0.2 mL/min. Calibration curves were achieved using standard solutions of related carboxylic acids. The identification of the carboxylic acids was performed by the retention time (t R ) comparison and standard addition methods using standard substances.

The TOC of the samples withdrawn from the treated solution at different electrolysis times was determined by Shimatzu TOC-V CSH analyser consisting of a non dispersive infra-red absorption detector (NDIR) according to the 680 C combustion catalytic oxidation method. Platinum based catalyst was used to facilitate the combustion at 650 °C. The carrier gas was oxygen with a flow rate of 150 mL/min. The injection volume was 50 μL.

Toxicity measurements

The toxicity of SAC and its intermediates generated in electro-Fenton processes was investigated on samples collected from solutions at different electrolysis times.

Experiments were performed with the bio-luminescence marine bacteria V. fischeri (Hach Lange France SAS) by Microtox method according to the international standard process (OIN 11348-3). Two values of the inhibition of the luminescence (%) were measured after 5 min and 15 min of exposure to samples at 15 C using a Microtox model 500 system.

Results and discussion

Effect of supporting electrolyte on the degradation of SAC

In order to investigate the effect of the supporting electrolyte on the degradation kinetics of SAC aqueous solutions, experiments were performed in acidic medium (pH 3.0) containing different supporting electrolytes as 50 mM Na 2 SO 4 , 100 mM NaNO 3 and 100 mM NaCl with three anodes (Fig. 4-1). The removal of SAC followed the first-order kinetic model and the apparent rate constant values (k app ) were determined For all the anodes, the complete removal of SAC almost finished in a 30 min electrolysis period in the presence of Na 2 SO 4 and NaNO 3 . However, it was removed only about 50% in 30 min in the case of NaCl. The main reason of this low degradation is related to the formation of active chlorine generated by the oxidation of chloride ions at the surface of DSA, Pt and BDD anode when NaCl used as supporting electrolyte (Eq.

(4-1)) (Özcan et al. 2008b;[START_REF] Loaiza-Ambuludi | Electro-Fenton degradation of anti-inflammatory drug ibuprofen in hydroorganic medium[END_REF][START_REF] De Moura | Active chlorine species electrogenerated on Ti/Ru 0.3 Ti 0.7 O 2 surface: Electrochemical behavior, concentration determination and their application[END_REF].

Therefore, part of electrical energy provided is lost in this reaction.

2Cl  → Cl 2 + e  (4-1)
The electrogenerated active chlorine can acts as oxidation mediator in the bulk of the solution, which can accelerate the removal rate of organic pollutants through the production of HClO (Eq. (4-2)) according to some researches (Özcan et al. 2008b;[START_REF] Loaiza-Ambuludi | Electro-Fenton degradation of anti-inflammatory drug ibuprofen in hydroorganic medium[END_REF].

Cl 2 + H 2 O → HClO + Cl  + H + (4-2)
However, in electro-Fenton process, the electrogenerated chlorine can also react with Fe 2+ (Eq. ( 4-3)) or decompose hydrogen peroxide (Eq. (4-4)) (Loaiza-Ambuludi et al.

2013), and then reducing the production rate of strong oxidant

Cl  + • OH → ClOH • (4-5)
Therefore, the removal efficiency of SAC was much lower when NaCl used as background electrolyte.

It can also be seen from Fig. 4-1 and Table 4-1 that the removal rate of SAC for Na 2 SO 4 was a little higher than that obtained for NaNO 3 in three electrolysis cells. For example, when DSA was used as anode and at 200 mA applied current, the k app for Na 2 SO 4 was 0.181 min 1 , while it was 0.170 min 1 for NaNO 3 . Therefore, Na 2 SO 4 was chosen to be the supporting electrolyte in the following experiments.

Effect of anode materials on the removal and mineralization of SAC

The removal of SAC by electro-Fenton process using DSA, Pt and BDD as anodes was investigated by keeping the same cathode as carbon-felt. Experiments were performed at Fe 2+ concentration 0.2 mM, Na 2 SO 4 concentration 50 mM, applied current 200 mA and pH 0 3.0. 0.2 mM SAC could be completely removed in a 30 min reaction for all the anode materials (Fig. 4-2a). The k app values for SAC removal using DSA, Pt and BDD anodes were very similar (0.181, 0.194 and 0.211 min 1 , respectively).

However, when it comes to mineralization, BDD anode showed its great superiority. In a 360 min reaction, the TOC removal efficiencies for SAC were 55.8%, 76.1% and 96.2%

for DSA, Pt and BDD anodes, respectively (Fig. 4-2b). On the one hand, the BDD(

Effect of Fe 2+ concentration on the removal of SAC

The effect of Fe 2+ (catalyst) concentration on the removal of SAC was examined by using the Fe 2+ concentration of 0.05, 0.1, 0.2, 0.3 and 0.5 mM. The time-course of SAC concentration decay was determined by HPLC, where SAC displayed a well-defined peak at the retention time (t R ) 14.2 min under the operating conditions of Fig. 4-3.

As can be observed in Fig. 4-3 and Table 4-1, the removal rate for SAC was enhanced by increasing the Fe 2+ concentration from 0.05 to 0.2 mM for both Pt and BDD anode. Moreover, the kinetics of SAC removal increases more rapidly (1.8 times) in the case of Pt/carbon-felt cell when the Fe 2+ increased from 0.05 to 0.2 mM, compared to BDD/carbon-felt cell. Since the applied current was kept constant at 200 mA, it can be assumed that the production rate of H 2 O 2 via the oxygen reduction would be identical for all the Fe 2+ concentrations under this study (Oturan et al. 2010). Then increasing the Fe 2+ concentration could lead to the increase of hydroxyl radicals generated via Fenton's reaction. These hydroxyl radicals reacted with SAC immediately, resulting in the increase of SAC degradation. On the contrary, further increasing the Fe 2+ concentration to 0.5 mM, the removal rate declined. The observed decrease in SAC degradation rate upon increasing the Fe 2+ concentration might due to the role of Fe 2+ as scavenger of hydroxyl radicals (Eq. (4-7)) which takes place with a large rate constant (k = 3.20 × 10 8 M 1 s 1 ) (Oturan et al. 2010). Therefore, the optimal Fe 2+ concentration for the removal of SAC was 0.2 mM.

Effect of applied current on the removal of SAC

The applied current is another important parameter for the effectiveness of the electro-Fenton process. In order to investigate the effect of current intensity on the oxidative degradation of SAC, several experiments were performed by varying applied current in the range of 50-500 mA at the optimal Fe 2+ concentration of 0.2 mM as assessed above. The results were shown in Fig. 4-4 and Table 4-1. When Pt was used as anode, the removal rate of SAC increased as the applied current increasing. The electrolysis time for complete disappearance of SAC has been changed from >> 30 min for 50 mA to 30, 25, 20 and 15 min for 100, 200, 300 and 500 mA current intensity, respectively (Fig. 4-4a). The k app was increased from 0.07 to 0.26 min 1 when applied current increased from 50 to 500 mA. The oxidation of SAC was accelerated by increasing the applied current because of progressively large production of In addition, the progressive enhancement of other parasitic reactions at higher current, such as hydrogen evolution (Eq. (4-10)) at the cathode and oxygen evolution (Eq. (4-11)) at the anode [START_REF] Yahya | Oxidative degradation study on antimicrobial agent ciprofloxacin by electro-fenton process: Kinetics and oxidation products[END_REF], also contributed to the lower SAC removal rate. It could be concluded that the comparatively better performance of BDD was more pronounced at lower current intensities.

2H 2 O + 4e  → H 2 + 2OH  (4-10) 2H 2 O → O 2 + 4H + + 4e  (4-11) (a) (b)

Determination of the rate constant of reaction between SAC and • OH

The absolute rate constant for the second order kinetics of the reaction between SAC and • OH (k abs,SAC ) was determined by using the competitive kinetics method [START_REF] Mhemdi | Electrochemical advanced oxidation of 2-chlorobenzoic acid using BDD or Pt anode and carbon felt cathode[END_REF]. Benzoic acid (BA) was employed as the standard competitor because the absolute rate constant of the reaction between BA and • OH (k abs,BA ) is known as 4.30 × 10 9 M -1 s -1 (Özcan et al. 2008a;Oturan et al. 2010). Experiments of competitive kinetics were carried out in the presence of equal concentrations of SAC and BA (0.1 mM). The hydroxylation rate constant of SAC can be calculated according to Eq. (4-12) and determined from Fig. 4-5 to be k abs,SAC = (1.85 ± 0.01) × 10 9 M -1 s -1 .

Ln (

[SAC] 0 [SAC] t ) = ( k abs,SAC k abs,BA ) Ln ( [BA] 0 [BA] t ) (4-12)
Interestingly, this rate constant value is the same as that reported by Toth et al (Toth et al. 2012). The effect of applied current on the mineralization of 0.2 mM SAC solution was investigated under the same condition as Fig. 4-4 and the results were shown in Fig. 4567.

When applied current raised from 50 to 200 mA, the TOC removal efficiency was increased from 64.2% to 76.1% in Pt/carbon-felt cell and from 88.9% to 96.2% in BDD/carbon-felt cell. However, the TOC removal efficiency decreased slightly for both anodes when current intensity further increased to 500 mA. As applied current increased, the higher electrogenerated H 2 O 2 concentration was obtained and larger amounts of • OH was yielded from Fenton's reaction, and then a greater mineralization efficiency can be achieved due to the simultaneous degradation of SAC and its byproducts (Dirany et al. 2012). But a current intensity higher than 200 mA would increase the extent of parasitic reactions (Eqs. [START_REF] Heng | Electro-Fenton removal of Orange II in a divided cell: Reaction mechanism, degradation pathway and toxicity evolution[END_REF][START_REF] Heng | Degradation of Orange II in aqueous solution by a novel electro/Fe 3 O 4 process[END_REF][START_REF] Heng Lin | Treatment of organic pollutants using electro-Fenton and electro-Fenton-like process in aqueous solution[END_REF](7)(8), [START_REF] Heng | Electro-Fenton removal of Orange II in a divided cell: Reaction mechanism, degradation pathway and toxicity evolution[END_REF][START_REF] Heng | Degradation of Orange II in aqueous solution by a novel electro/Fe 3 O 4 process[END_REF][START_REF] Heng Lin | Treatment of organic pollutants using electro-Fenton and electro-Fenton-like process in aqueous solution[END_REF](7)(8)(9)(10) and [START_REF] Heng | Electro-Fenton removal of Orange II in a divided cell: Reaction mechanism, degradation pathway and toxicity evolution[END_REF][START_REF] Heng | Degradation of Orange II in aqueous solution by a novel electro/Fe 3 O 4 process[END_REF][START_REF] Heng Lin | Treatment of organic pollutants using electro-Fenton and electro-Fenton-like process in aqueous solution[END_REF](7)(8)(9)(10)(11)) and decrease the mineralization efficiency of SAC (Mousset et al. 2014). when initial SAC concentration was 0.2 mM, Fe 2+ concentration was 0.2 mM, Na 2 SO 4 concentration was 50 mM, applied current was 200 mA and pH 0 was 3.0.

Oxalic, formic, and maleic acid at retention time (t R ) of 8.90, 16.03 and 11.33, respectively, were observed during electro-Fenton process. Since the concentration of maleic acid was detected in trace level, the changes of maleic acid concentration were not shown in Fig. 45678. In Pt/carbon-felt cells, formic acid presented only in the first 20 min and then it undergo to the mineralization (Fig. 4-8a) (Özcan et al. 2008a). Oxalic acid showed the largest accumulation and the higher persistence due to its lower reactivity with • OH, which suggested that it can be generated from different ways as ultimate reaction intermediate (Dirany et al. 2010). In addition, oxalic acid was still existed with a concentration of 0.04 mM after a 360 min reaction. This relatively weak degradation of oxalic acid can be attributed to its high stability in the presence of ferric ions and its weaker reactivity toward • OH radicals (k 2 = 2.8 × 10 7 M -1 s -1 ) (Oturan et al. 2008). This was also corresponding to the residual TOC remaining at the end of mineralization treatments in Pt/carbon-felt cells.

On the other hand, oxalic acid showed high persistence, but it disappeared at the end of the reaction in BDD/carbon-felt cell (Fig. 4-8b). This phenomenon was in agreement with the faster TOC removal in BDD/carbon-felt cells as described in section 4.3.6.

Evolution of SAC solution toxicity during electro-Fenton process

In order to determine the potential toxicity of SAC and its intermediates,

Conclusions

It was demonstrated that electro-Fenton process with a Pt or BDD anode and a carbon felt cathode was a very effective method for the degradation of SAC in water. The effect of Fe 2+ concentration and applied current on the removal and mineralization of SAC was investigated. For both anodes, SAC could be completely degrade in less than 30 min. The oxidative degradation of SAC followed pseudo-first-order kinetics.

Absolute rate constant of hydroxylation reaction of SAC was determined as (1.85 ± 0.01) × 10 9 M -1 s -1 using competition kinetics method. The optimal conditions for SAC removal were SAC concentration 0.2 mM, Fe 2+ concentration 0.2 mM, Na 2 SO 4 concentration 50 mM, applied current 200 mA and initial pH 3.0. The use of BDD anode yields a faster mineralization rate because of the higher oxidation power of BDD comparing with Pt. Short-chain aliphatic carboxylic acids such as oxalic, formic and maleic acids are identified as aliphatic by-products. The bacteria luminescence inhibition showed the toxicity of SAC solution increased at the first 30 min of electrolysis, and then it decreased after a 60 min reaction for both Pt and BDD anodes. in aqueous medium by anodic oxidation using boron-doped diamond anode: Influence of experimental parameters on degradation kinetics and mineralization efficiency. Water Research 42(12), 2889-2898.

Siré s, I., Garrido, J.A., Rodrí guez, R.M., Brillas, E., Oturan, N. and Oturan, M.A. (2007) Catalytic behavior of the Fe 3+ /Fe 2+ system in the electro-Fenton degradation of the antimicrobial chlorophene. Applied Catalysis B: Environmental 72 [START_REF] Heng | Degradation of bisphenol A in aqueous solution by a novel electro/Fe 3+ /peroxydisulfate process[END_REF][START_REF] Heng | Electro-Fenton removal of Orange II in a divided cell: Reaction mechanism, degradation pathway and toxicity evolution[END_REF], 382-394. Toth, J.E., Rickman, K.A., Venter, A.R., Kiddle, J.J. and Mezyk, S.P. ( 2012 Chemosphere 117, 447-454.

Introduction

In this part, the effect of the Fe 2+ concentration and applied current on the mineralization of sucralose (SUC) was examined. The effect of applied current on the mineralization efficiency was assessed. The aliphatic short-chain carboxylic acids and inorganic ions released during electro-Fenton process were monitored by ion-exclusion chromatography and ion chromatography (IC). The variation of toxicity of SUC solution and its intermediates was determined by Microtox method.

Materials and methods

Chemicals

Sucralose (1,6-dichloro-1,6-dideoxy-β-D-fructofuranosyl-4-chloro-4- Analytical grade carboxylic acids and other chemicals used for chromatographic analysis were purchased from Acros, Merck, Sigma, Riedel-de Haë n and Fluka.

Ultrapure water used for the preparation of the working solutions and HPLC eluting solutions was obtained from a Millipore Milli-Q (simplicity 185) system with resistivity > 18 MΩ cm at room temperature.

Electrochemical apparatus and procedures

Bulk experiments were conducted in an open, undivided and cylindrical glass cell of 250 mL capacity containing 220 mL SUC solution. Either a cylindrical Pt mesh (4.5 cm height, i.d. = 3.1 cm, Platecxis, France) or a 24 cm 2 thin-film BDD electrode (CONDIAS GmbH, Germany) was used as anode, and a 87.5 cm 2 piece of carbon felt (17.5 × 5 cm, Carbon-Lorraine, France) was used as cathode. In all electrolyses, the anode was centered in the cell, surrounded by the carbon felt which placed on the inner wall of the cell covering the totality of the internal perimeter. H 2 O 2 was generated from the reduction of O 2 dissolved in the solution. The continuous saturation of oxygen at atmospheric pressure was assured by bubbling compressed air through a frit at about 0.5 L/min, starting 5 min before electrolysis. All the electrolytic trials were conducted under constant current conditions by a Hameg HM8040-3 triple power supply (Germany).

The degradation experiments were performed using 0.2 mM SUC solution with constant stirring by using a magnetic stirrer. 50 mM Na 2 SO 4 was added into the SUC solution as supporting electrolyte. A catalytic quantity of ferrous ion was added into the solution before the beginning of electrolysis. The initial pH (pH 0 ) of SUC solutions was measured with a CyberScan pH 1500 pH-meter (Eutech Instrument, USA) and set at 3.0 (±0.1), adjusting by the addition of 1 M sulfuric acid.

Analytical methods and procedures

The TOC of the samples withdrawn from treated solutions at different electrolysis times were determined by Shimatzu TOC-V CSH analyser consisting of a non-dispersive infra-red absorption detector (NDIR) according to the 680 C combustion catalytic oxidation method. Platinum was applied as catalyst to facilitate the combustion at 650 °C. The carrier gas was oxygen with a flow rate of 150 mL/min. The injection volume was 50 μL.

The short-chain carboxylic acids were identified and quantified by ion-exclusion HPLC using a L-7100 pump, a Merck Lachrom liquid chromatograph equipped with a Supelcogel H column (250 × 4.6 mm, 9 μm) and a L-7455 photodiode array detector at the wavelength of 220 nm. The mobile phase was 0.1% H 2 SO 4 solution and flow rate was fixed to 0.2 mL/min. Calibration curves were achieved using standard solutions of related carboxylic acids. The identification of the carboxylic acids was performed by the retention time (t R ) comparison and standard addition methods using standard substances.

Inorganic ions (Cl -) released in the electro-Fenton process were monitored by ion chromatography with a Dionex ICS-1000 Basic Ion Chromatography (IC) System equipped with an IonPac AS4A-SC (anion exchange) 250 × 4 mm column and fitted to a DS6 conductivity detector containing a cell heated at 35 C under control through a Chromeleon SE software. The mobile phase was a mixture of 3.6 mM Na 2 CO 3 and 3.4 mM NaHCO 3 solution with a flow rate of 2.0 mL/min. The volume of injections was 25 μL.

Toxicity measurements

The toxicity of SUC and its intermediates generated during the electrolyses were investigated via samples collected from treated solutions at different times. Experiments were performed with the bio-luminescence marine bacteria V. fischeri (Hach Lange France SAS) by Microtox method according to the international standard process (OIN 11348-3). Two values of the inhibition of the luminescence (%) were calculated after 5 and 15 min of exposure to samples at 15 C.

Results and discussion

Effect of Fe 2+ concentration on the mineralization of sucralose

Mineralization efficiency in both Pt/carbon-felt and BDD/carbon-felt cells was assessed by solution TOC (as equivalent of dissolved organic carbon) measurements during electrolysis of SUC solutions. The effect of Fe 2+ concentration on the mineralization of SUC solution was monitored by using the Fe 2+ concentration of 0.05, 0.1, 0.2, 0.3 and 0.5 mM. The SUC concentration was fixed at 0.2 mM, applied current was 200 mA and initial pH was 3.0. As can be observed in Fig. 5-1, the mineralization for SUC was enhanced by increasing the Fe 2+ concentration from 0.05 to 0.2 mM for both Pt and BDD anode. Since the applied current was kept constant at 200 mA, it can be assumed that the production rate of H 2 O 2 via the oxygen reduction would be identical for all the Fe 2+ concentrations under this study (Oturan et al. 2010). On the basis of the same H 2 O 2 production rate, a higher Fe 2+ concentration could lead to the increase of hydroxyl radicals generated in bulk of solution via Fenton's reaction (Eq. (5-1)). These hydroxyl radicals would react with SUC immediately, resulting in the increase of SUC mineralization.

Fe 2+ + H 2 O 2 → Fe 3+ + OH  + • OH (5-1)
However, further increasing the Fe 2+ concentration to 0.5 mM, the mineralization efficiency of SUC declined evidently in both Pt/carbon-felt and BDD/carbon-felt cells.

The negative influence of higher Fe 2+ concentration might due to the role of Fe 2+ as scavenger of hydroxyl radicals (Eq. (5-2)) which takes place with a large rate constant s et al. 2007b;Oturan et al. 2010). Therefore, the value of 0.2 mM was chosen as the optimal Fe 2+ concentration under these conditions and used in the following experiments.

(k = 3.20 × 10 8 M 1 s 1 ) (Siré
Fe 2+ + • OH → Fe 3+ + OH  (5-2)
By comparing Fig. 5-1a and Fig. 5-1b, it can be seen that the use of BDD anode lead to the acceleration of SUC mineralization. At optimal condition (Fe 2+ concentration 0.2 mM), 96.1% of the initial TOC value was removed in BDD/carbon-felt cell after only 180 min reaction, while it takes 240 min for Pt/carbon-felt cell to achieve 94.2% of mineralization efficiency. This fact can be explained by following reasons: Firstly, a much higher quantities of BDD( • OH) than Pt( • OH) is provided when electrolysis is operated at the current within the water discharge region (Eq. ( 5-3) and (5-4)) (Brillas et al. 2005;Panizza and Cerisola 2005;Brillas et al. 2009).

BDD + H 2 O → BDD( • OH) + H + + e  (5-3) Pt + H 2 O → Pt( • OH) + H + + e  (5-4)
Secondly, since BDD anode presents the highest O 2 -overvoltage (1.27 V), which is much higher than Pt anode (0.27 V) (Martí nez-Huitle and Brillas 2009; Oturan et al. 2011), the oxidative action of BDD( • OH) is rather effective than Pt( • OH) (Oturan et al. 2011). Thirdly, due to the low adsorption ability of • OH on BDD, the physisorbed and labile BDD( • OH) formed at the anode surface (Eq. (5-3)) can readily react with organic pollutant, in contrast to the chemisorbed radicals typically formed at Pt surface (Eq.

(5-4)) which limited the oxidation ability of Pt( • OH) (Oturan et al. 2011;Dirany et al. the most refractory ones (Dirany et al. 2012). In order to investigate the effect of current intensity on the mineralization of SUC, several experiments were carried out by varying applied current in the range of 100-500 mA at the optimal Fe 2+ concentration of 0.2 mM as assessed above. The results were shown in Fig. 5-2. In both Pt/carbon-felt and BDD/carbon-felt cells, the mineralization efficiency of SUC solution increased significantly as the applied current increased from 100 to 200 mA. Taking Pt/carbon-felt cell as an example, when current intensity was 100 mA, the TOC removal efficiency was 69.7% in a 360 min treatment. Increasing the current to 200 mA resulted in 94.1% of mineralization efficiency in a 240 min reaction.

The mineralization of SUC aqueous solution was accelerated by increasing the applied current because both Fe 2+ regeneration and H 2 O 2 production would be promoted, and then large amount of • OH would be produced (Boye et al. 2002;Lin et al. 2014). When the applied current increased from 200 to 500 mA, the minralization efficiency of sucralose in both cells changed slightly. On the one hand, a higher current leads to the increase of parasitic reactions, such as hydrogen evolution (Eq. ) and oxygen evolution (Eq. (5-6)) reactions (Dirany et al. 2010;Dirany et al. 2012).

2H 2 O + 2e -→ H 2 + 2OH - (5-5) 2H 2 O → O 2 + 4H + + 4e  (5-6)
On the other hand, when the current density was further increased, excessive hydroxyl radicals would be consumed via the following side reactions (Eqs. (5-7) and

(5-8)) (Wu et al. 2012;Lin et al. 2014). As a result, 200 mA was selected as the optimal current value and used in the remaining experiments.

H 2 O 2 + • OH → HO 2 • + H 2 O (5-7) HO 2 • + • OH → H 2 O + O 2 (5-8)
It also be observed from

Mineralization current efficiency (MCE)

The mineralization capacity of the process can be expressed by MCE which can be determined from the following equation (Eq. (5-9)) [START_REF] Skoumal | Mineralization of the biocide chloroxylenol by electrochemical advanced oxidation processes[END_REF]Brillas et al. 2009;Oturan et al. 2011

): MCE = s C (5-9)
where n is the number of electrons consumed per SUC molecule according to Eq. (5-10) (= 48), F is the Faraday constant (= 96,487 C/mol), V s is the solution volume (L), Δ (TOC) exp is the experimental TOC decay (mg/L), 4.32 × 10 7 is the conversion factor to homogenize units (= 3600 s/h × 12,000 mg of C/mol), m is the number of carbon atoms of sucralose (= 12), I is the applied current (A) and t is the given treatment time

(h). C 12 H 19 Cl 3 O 8 + 16H 2 O → 12CO 2 + 51H + + 3Cl  + 48e 
(5-10) By comparing Fig. 5-3(a) and Fig. 5-3(b), it can be seen that the MCE in Pt/carbon-felt cell was lower at all the chosen time than that in BDD/carbon-felt cell.

This was attributed to the low oxidation power of Pt anode (Oturan et al. 2011).

Figure 5-3 indicated that the MCE values decreased with rising applied current from 100 to 500 mA in both Pt/carbon-felt and BDD/carbon-felt cells. For example, the MCE values after 60 min of electrolysis in BDD/carbon-felt cell for 100, 200, 300 and 500 mA were 31.49%, 20.17%, 14.40% and 8.90%, respectively. This was related to the involvement of some parasitic reactions and particularly the hydrogen gas evolution reaction (Eq. (5-5)), which competed with the formation of H 2 O 2 (Eq.(5-11)) (Özcan et al. 2008).

O 2 + 2H + + 2e  → H 2 O 2
(5-11)

Figure 5-3 also showed that the MCE values decreased continuously from the beginning to the end of the electrolysis for every trial in both cells. This was due to the gradual formation of intermediates such as carboxylic acids that are more difficult to destroy by BDD( • OH) and also due to mass transport limitations because of the presence of small concentration of organic matter (Siré s et al. 2007a;Özcan et al. 2008).

Identification and evolution of short-chain carboxylic acids

To identify the generated carboxylic acids released in the electro-Fenton process, 0.2 mM SUC solution aqueous solution was treated when Fe 2+ concentration was 0.2 mM, Na 2 SO 4 concentration was 50 mM, applied current was 200 mA and pH 0 was 3.0.

The concentration of carboxylic acids was monitored by ion-exclusion HPLC and the results were shown in Fig. 5-4. In both Pt/carbon-felt and BDD/carbon-felt cells, oxalic, pyruvic, formic and glycolic acid at retention time (t R ) 8.90, 11.57, 16.03 and 14.53, respectively, were observed. As shown in Fig. 5-4, formic acid was observed as soon as the electrolysis started in both cells. Its concentration reaches the maximum value of 0.090 mM after a 20 min reaction in Pt/carbon-felt cell and 0.096 mM after only 10 min in BDD/carbon-felt cell. Then it decreased gradually to be completely mineralized after 180 min and 60 min in Pt/carbon-felt and BDD/carbon-felt cells, respectively. Pyruvic acid was also generated at the beginning of the electrolysis and followed the accumulation-destruction cycles in two cells. It disappeared after 120 min electrolysis in both cells. The accumulation of glycolic acid was a little slower and appeared after 7 min of electrolysis. Oxalic acid showed the largest accumulation and the higher persistence in both cells due to its high stability in the presence of ferric ions and its weaker reactivity toward • OH radicals (k 2 = 2.8×10 7 M -1 s -1 ) (Oturan et al. 2008).

This suggested that oxalic acid would be generated from different ways as ultimate reaction intermediates.

Moreover, all the carboxylic acids were vanished after a 360 min electrolysis in Pt/carbon-felt cell and a 240 min electrolysis in BDD/carbon-felt cell, which was corresponding to the residual TOC remaining at the end of mineralization treatments (Fig. 5-2a and Fig. 5-2b). toxic than the previous sucralose solution in the first 60 min electrolysis [START_REF] Feng | Degradation of anti-inflammatory drug ketoprofen by electro-oxidation: comparison of electro-Fenton and anodic oxidation processes[END_REF]. The other peak appeared at 120 min, indicating the generation of other toxic byproducts. From 120 to 240 min, the inhibition percentage decreased sharply. The toxicity was much lower than the untreated sucralose solution at 240 min, in relation to the increase of the less toxic byproducts, such as carboxylic acids, which present a very low toxicity toward V. fischeri bacteria (Dirany et al. 2012). Consequently, the toxicity kept stable from 240 min to the end of the electrolysis.

The curves obtained for BDD anode showed a similar behavior at the beginning of the electrolysis as Pt anode. Several peaks were obtained, indicating the formation of some intermediates that were more toxic than the previous sucralose solution in a 60 min reaction. The secondary peak appeared at 40 min and showed the secondary intermediates still more toxic than the original sucralose solution. The secondary peak which appeared at 240 min was much lower than the primary peak, showing the toxicity of secondary intermediates were much lower than the primary byproducts. After 240 min, the inhibition percentage decreased significantly and was much lower than the untreated sucralose solution at the end of the electrolysis, showing the disappearance of toxic intermediate products.

Conclusions

Electro-Fenton process using a carbon-felt cathode and a Pt or BDD anode was successfully applied to the mineralization of sucralose in aqueous solutions in this study.

The optimal conditions for the mineralization of 0.2 mM sucralose solution were Fe 2+ concentration 0.2 mM, Na 2 SO 4 concentration 50 mM, applied current 200 mA and initial pH 3.0. At optimal conditions, sucralose could be totally mineralized in 240 min for BDD anode and 360 min for Pt anode. Short-chain aliphatic carboxylic acids such as oxalic, pyruvic, formic and glycolic are identified as aliphatic byproducts during the electro-Fenton process. Cl atoms were found in the structure of sucralose and were converted to Cl  . The bacteria luminescence inhibition showed the toxicity of sucralose solution increased at the beginning of electrolysis, and then it declined until much lower than the original sucralose solution at the end of the reaction.

Introduction

In this part, peroxydisulfate (PDS), which has the similar structure of O-O bond contained in H 2 O 2 [START_REF] Hou | Ultrasound enhanced heterogeneous activation of peroxydisulfate by magnetite catalyst for the degradation of tetracycline in water[END_REF], was activated by goethite to generate sulfate radicals and remove Orange II from water in a electrochemical (EC) reactor (electro/α-FeOOH/peroxydisulfate process).

The traditional one-factor-at-a-time approach has been widely used to study the effects of various factors in most experiments. Experimental factors are varied one at a time, with the remaining factors being held constantly [START_REF] Zhang | Application of experimental design methodology to the decolorization of Orange II using low iron concentration of photoelectro-Fenton process[END_REF]. This method fails to investigate the interactive influences of different variables. When a combination of several independent variables and their interactions affect decolorization efficiency, response surface methodology (RSM) is effective for optimizing the operating parameters in multivariable systems [START_REF] Gong | Evaluation of heterogeneous photo-Fenton oxidation of Orange II using response surface methodology[END_REF][START_REF] Zhang | Application of experimental design methodology to the decolorization of Orange II using low iron concentration of photoelectro-Fenton process[END_REF]. As one of the RSM designs, Box-Behnken design (BBD) is known as a modified central composite experimental design [START_REF] Ay | A statistical experiment design approach for advanced oxidation of Direct Red azo-dye by photo-Fenton treatment[END_REF][START_REF] Zhang | Application of experimental design methodology to the decolorization of Orange II using low iron concentration of photoelectro-Fenton process[END_REF]. BBD is an independent, rotatable quadratic design with no embedded factorial or fractional factorial points where the variable combinations are at the midpoints of the edges of the variable space and at the center. In addition, BBD requires fewer runs than other RSM designs with the same number of variables. For example, only 15 runs are needed for a three-parameter experimental design. As a result, BBD was applied in this study to investigate the effect of electric current, α-FeOOH concentration and PDS concentration on the results of the pseudo-first-order decolorization rate constant of Orange II (a kind of widely used zao dyes) in electro/α-FeOOH/peroxydisulfate process.

Materials and methods

Chemicals

Orange II (4-(2-hydroxynaphthylazo) benzenesulfonic acid sodium salt) was purchased from Sinopharm Chemical Reagent Co., Ltd. (China) and used without further purification. The Chemical structure and main characteristics of Orange II were presented in Table 6-1. α-FeOOH was obtained from Sigma-Aldrich. Stock solution was prepared by dissolving a selected amount of Orange II in deionized water and the initial concentration (C 0 ) was fixed at 50 mg/L. 50 mM Sodium sulfate was added as a supporting electrolyte. Motors Factory, China) was applied to achieve the transport of electroactive matter toward the electrode and ensure homogeneity throughout the reaction. The DC power supply was initiated when PDS solutions and α-FeOOH were added to the electrolytic cell. Samples were taken from electrolytic reactor at each predetermined time interval.

These samples were filtered through 0.22 μm membranes (Millipore Co.) before analysis and then mixed with the same volume of methanol to quench the reaction before analysis.

Analytical methods and procedures

The absorbence of Orange II was measured at maximum absorption wavelength of 485 nm using a Rayleigh UV-9100 spectrophotometer (Rayleigh Co., China). The UV-visible spectrum from 200 to 800 nm was achieved by a spectrophotometer (Shimadzu, UV-1700). The concentration of residual PDS was measured by an iodometric titration method (Wahba et al. 1959).Total organic carbon (TOC) analyses were carried out with non dispersive infra red absorption detector (NDIR) using an Analytikjena multi N/C 3100 analyzer. The carrier gas was oxygen with a flow rate of 150 mL/min. The structure of α-FeOOH was characterized by X-ray diffraction (XRD).

XRD patterns are collected on a X'Pert Pro Xray diffractometer (PANalytical B.V., Holland), using a Cu Kα radiation (λ = 1.5406 Å).The X-ray photoelectron spectroscopy (XPS) was operated on a ESCALAB 250Xi spectrometer (Thermo fisher, USA) which equipped with AlKα X-ray source. The XPS data analysis was performed using the XPSPeak4.1 program with a symmetric Gaussian-Lorentzian sum function and Shirley background subtraction.

Box-Behnken design (BBD)

Box-Behnken statistical experiment design and response surface methodology were employed to investigate the effects of the three independent variables on the response functions. The independent variables were electric current (X 1 ), α-FeOOH dosage (X 2 ) and PDS concentration (X 3 ). The low, center and high levels for each variable are designed as -1, 0 and +1, respectively as illustrated in Each independent variable was coded as X i and determined by the following equation:

0 i i i xx X x    (6-1)
where X i is dimensionless code value of the i th independent variable, x i is the uncoded value of the ith independent variable, x 0 is the uncoded i th independent variable at the center point and Δx i is the step change value between low level (-1) and high level (+1).

As can be seen in Table 6-3, the total number of experiments in this study was 15 based on 3 levels and a 3 factor experimental design, with three replicates at the center of the design for estimation of a pure error sum of squares. Experimental data from the BBD could be analyzed and fitted to a second-order polynomial model using Design Expert 8.0 software:

Y = β 0 + ∑β i X i + ∑β ii X 2 i + ∑β ij X i X j (6-2)
where Y is the response and X i and X j are the independent variables, square effects and interaction effects; β i , β ij and β ii are the linear coefficients, interaction coefficients and squared coefficients; β 0 is the intercept parameter. ( 9)) (Özcan et al. 2009).

M(H 2 O) → M( • OH) + H + + e - Adding 0.5 g/L α-FeOOH to the electrolytic reactor (EC/α-FeOOH process) could not improve the decolorization efficiency of Orange II. The decolorization efficiency was 57.2% after 120 min reaction in EC/α-FeOOH process. The decolorization efficiency improved significantly to 81.4% in EC/PDS process, which might be attributed to sulfate radicals produced via Eq. . 50.0% of PDS was decomposed in this case (Fig. 6-1b).

S 2 O 8 2-+ e -→ SO Based on the results and the previous studies, (Liang et al. 2007;Özcan et al. 2009;Lin et al. 2013), the mechanism of EC/α-FeOOH/PDS process was shown in Fig. 6-4.

Fe(II) was generated by the reduction and transformation of Fe(III) on the surface of α-FeOOH due to the direct electron transfer (Eq. ), and then Fe(II) activated PDS to generate sulfate radicals (Eq. ). Meanwhile, hydroxyl radicals were generated both by water discharge on DSA anode (Eq. ) and the reaction between sulfate radicals and H 2 O or OH -(Eqs. (6-7) and (6-8)) (Liang et al. 2007;Saien et al. 2011).

Orange II was degraded by both sulfate radicals and hydroxyl radicals.

Effect of initial pH

The effect of pH on degradation of Orange II was investigated when initial Orange II concentration was 50 mg/L, α-FeOOH dosage was 0.5 g/L, PDS concentration was 2 g/L, Na 2 SO 4 solution concentration was 50 mM and the electrical current was 0.5 A.

The initial pH values investigated were 3, 7 and 10. As it indicated in Fig. 6-5, no significant effect of pH on the decolorization efficiency was observed. The final efficiency was 90.9%, 92.1% and 89.1% at initial pH 3.0, 7.0 and 10.0, respectively. It is reported that sulfate radicals could react with H 2 O or OH -to produce hydroxyl radicals (Eqs. [START_REF] Heng Lin | Treatment of organic pollutants using electro-Fenton and electro-Fenton-like process in aqueous solution[END_REF](7) and (6-8)) (Liang et al. 2007;Saien et al. 2011). Either H + could be generated or OH -could be consumed, which led to the reduction of solution pH in EC/FeOOH/PDS process. Moreover, the reaction of hydroxyl radical and sulfate radical 123 ) as well as the decomposition of PDS in water (Eq. ( 6 Therefore, the final pH was almost the same value (about 2.5) and the decolorization of Orange II tended to be operated under similar conditions as the reaction proceeded. This indicated that the EC/α-FeOOH/PDS process could be successfully applied at a wide range of initial pH. Since pH of most actual texile effluent was neutral, pH 7.0 was chosen as the optimal initial value used in the following experiments.

Response surface analysis

Main variables

The full factorial BBD with three factors in three levels as well as results of Orange II decolorization rate constant for each run are listed in Table 6-3. Based on Table 6-3, the main effects plot and the interaction plots for Orange II decolorization rate constant at 120 min were developed. The main effects plot illustrated that the effects of electrical current, α-FeOOH dosage and PDS concentration on Orange II decolorization rate constant were positive (Fig. . This means that the greater removal efficiency could be achieved at high level (+1) of each factor than that at low level (-1) of the factor. Among the variables, the influence of electrical current was the greatest, and the second one was the PDS concentration. The influence of α-FeOOH dosage was softer than the other two variables at 120 min. among the factors. One factor was fixed at its high (+1) or low (-1) level while the other was investigated [START_REF] Domínguez | Electrochemical advanced oxidation of carbamazepine on Boron-doped diamond anodes. Influence of Operating Variables[END_REF]. As can be seen in Fig. 6-4, the curves of interaction between α-FeOOH dosage and PDS concentration are across, indicating the interaction between two variables is significant. The curves of interaction between electrical current and PDS concentration are tending to cross. It may assume that the interaction between two variables is significant. The interaction effect between electrical current and α-FeOOH dosage are less significant than that between α-FeOOH dosage and PDS concentration. This was also confirmed by the high probability value ((Prob > F) > 0.1) through analysis of variance (ANOVA).

Fitting model and analysis of variance

In order to justify the significance and adequacy of the model, the ANOVA analysis is required. The value of (Prob > F) was used as a tool to check the significant of each model terms. The smaller the value of (Prob > F) means the more significant of the corresponding model term. Generally, the value of (Prob > F) over 0.1 implies that the model term is insignificant [START_REF] Zhang | Application of response surface methodology to the treatment landfill leachate in a three-dimensional electrochemical reactor[END_REF]. Based on the results shown in Table 6-4, RSM model of second order polynomial equations was determined as follows, would give poor or misleading results if the model was not adequate fit (Wu et al. 2012a). The normal probability and studentized residuals plots illustrate whether the studentized residuals follow a normal distribution. Normal plot presented at Fig. 6-8 is normally distributed and resemble a straight line, indicating that there is no apparent problem with normality and no need for transformation of response [START_REF] Khataee | Optimization of photocatalytic treatment of dye solution on supported TiO 2 nanoparticles by central composite design: Intermediates identification[END_REF]. Fig. 6-9 shows the studentized residual versus predicted value, and the residuals appear to be a random scatter. It suggests that the equality of variance does not seem to be violated [START_REF] Grčić | Minimization of organic content in simulated industrial wastewater by Fenton type processes: A case study[END_REF]. Furthermore, the similarity between predicted and actual values of the response which was illustrated in Fig. 6-10 indicated that there was no significant violation of the models and the models were satisfactory and accurate.

Stability of α-FeOOH

The recycle experiments were performed to investigate the stability of α-FeOOH.

The solid was easily removed from the electrochemical reactor after each repetitive oxidation process, then washed by deionised water, dried in the vacuum oven and stored at ambient temperature. The initial concentration of Orange II solution was fixed at 50 mg/L, PDS concentration was 2 g/L, α-FeOOH dosage was 0.8 g/L, Na 2 SO 4 solution concentration was 50 mM, initial pH was 7.0 and the applied current was 0.5 A. As shown in Fig. 6-12, the decolorization efficiencies of Orange II during three reaction cycles ranged from 86.5% to 92.1%, indicating α-FeOOH is stable and can be reused. corresponding spectra are indicated in Fig. 6-13. As can be seen from these spectra, there was one main band located at 485 nm in the visible region, which was originated from the azo-band chromophore (N=N) [START_REF] Wu | Degradation of Acid Orange 7 in aqueous solution by a novel electro/Fe 2+ /peroxydisulfate process[END_REF]. Meanwhile, the absorption at 310 nm and 228 nm in the ultraviolet region were corresponded to naphthalene ring and benzene ring structures in the dye molecule [START_REF] Zhang | Degradation of C.I. Acid Orange 7 by ultrasound enhanced heterogeneous Fenton-like process[END_REF].

As shown in Fig. 6-13(a), the visible band disappeared as the reaction proceeded, due to the fragmentation of the azo links by the reaction. The decay of the absorbance at 310 nm was considered as evidence of aromatic fragment degradation in the dye molecule and its intermediates. However, the absorbance at 228 nm increased, indicating that the azo structure was destroyed and some aromatic fragment such as benzene rings were formed.

In order to investigate the mineralization efficiency of Orange II electro/α-FeOOH/peroxydisulfate process, the changes of TOC were observed with initial Orange II concentration 50 mg/L, α-FeOOH dosage 0.5 g/L, PDS concentration 2 g/L, Na 2 SO 4 solution concentration 50 mM, initial pH 7.0 and the electrical current 0.5A.

As can be seen in Fig. 6

Materials and methods

Materials

Orange II (4-(2-hydroxynaphthylazo)

Experimental

The experimental set-up used for this study is illustrated in Fig. 7-1. Bulk electrolyses were carried out in undivided glass beaker with a capacity of 250 mL. A plate mixed metal oxide (DSA, Ti/RuO 2 IrO 2 ) anode (5.0 cm × 11.9 cm) is served as the anode and a plate of stainless steel with the same dimensions is used as cathode. The gap between the electrodes was 4.0 cm. The reactor was immersed in a water bath to where the A 0 and A were the absorbance of the sample at time 0 and t, respectively. The concentration of residual PDS was measured by an iodometric titration method (Wahba et al. 1959). The quantification of the dissolved iron ions concentration in the solution

Decolorization of Orange II under different systems

To Based on the results and the literature (Liang et al. 2007;Özcan et al. 2009;Yan et al. 2011;Lin et al. 2013), the mechanism of EC/Fe 3 O 4 /PDS process was indicated in Fig. 7-4. Fe(II) was generated by the reduction and transformation of Fe(III) on the surface of Fe 3 O 4 due to the direct electron transfer (Eq. (7-6)), and then Fe(II) activated PDS to release sulfate radicals via Eq. (7-7)). Meanwhile, hydroxyl radicals were generated both by water discharge on DSA anode (Eq. (7-4)) and the reaction between sulfate radicals and H 2 O or OH -(Eqs. (7-8) The degradation kinetics of different systems was investigated and the results were summarized in

Effect of initial pH

The effect of initial pH was observed when Orange II concentration was 25 mg/L, PDS concentration was 20 mM, Fe 3 O 4 dosage was 0.5 g/L, Na 2 SO 4 solution concentration was 50 mM and the electrical density was 8.4 mA/cm 2 . As can be seen from Fig. 7-5, the decolorization efficiencies of Orange II were 93.0%, 91.2% and 92.1% at initial pH 3.0, 6.0 and 9.0, respectively. When initial pH ranged from 3.0 to 9.0, the rate constants varied slightly from 0.037 to 0.043 min -1 (Table 7-1). Sulfate radicals generated in EC/Fe 3 O 4 /PDS process could react with H 2 O or OH -to produce hydroxyl radicals (Liang et al. 2007;Saien et al. 2011). Either H + could be generated or OH - could be consumed (Eqs. (7-8) and (7-9)), which resulted in the reduction of solution pH. Moreover, the dissociation of HSO 4 -(Eq. (7-10)), which was formed through Eq.

(7-11), would further release H + and decrease the solution pH (Kusic et al. 2011). 7-1) in a 60 min reaction when current density was 3.4 mA/cm 2 . When current density increased to 25.2 mA/cm 2 , the decolorization efficiency was nearly 100% in 30 min reaction and the rate constant increased to 0.117 min -1 (Table 7-1). On the one hand, when current density increased, more hydroxyl radicals would be formed on anode surface (Eq. (7-4)) and more sulfate radicals would be generated via electron transfer reaction (Eq. (7-5)). On the other hand, the increase of current density favored the cathodic reduction of Fe(III) (Eq. (7-6)), and more PDS would be activated to generate sulfate radicals via Eq. (7-7). Meanwhile, did cause the increase of PDS decomposition. This also suggested more sulfate radicals would be generated when current density increased.

Effect of PDS concentration

As the source of sulfate radicals, PDS plays a significant role in EC/Fe 3 O 4 /PDS process. The effect of PDS concentration was investigated by varying the concentration of PDS from 5 to 20 mM when Orange II concentration was 25 mg/L, Fe 3 O 4 dosage was 0.5 g/L, Na 2 SO 4 solution concentration was 50 mM, initial pH was 6.0 and the electrical density was 8.4 mA/cm 2 . As can be seen from Fig. 7-7, when PDS concentration increased from 5 to 10 mM, the Orange II decolorization efficiency and apparent rate constants over a 60 min reaction increased from 84.1% to 95.8% and the rate constants increased from 0.030 to 0.053 min -1 (Table 7-1). Meanwhile, the amount of PDS decomposed also increased, indicating more reactive radicals would be generated to degrade Orange II at higher PDS concentration. When PDS concentration increased to 20 mM, the Orange II decolorization efficiency and rate constants after 60 min declined slightly to 91.2% and 0.039 min -1 (Table 7-1). This was due to the fact that the side reaction between S 2 O 8 2-and SO

Electrical energy consumption estimation

In order to evaluate the energy consumption of EC/Fe 3 O 4 /PDS process, an electrical energy consumption per order of magnitude (EE/O) was employed. EE/O is defined as the number of kilowatt hours of electrical energy required for bringing about the degradation of a pollutant by one order of magnitude (90%) in 1 m 3 of contaminated water [START_REF] Daneshvar | Removal of C.I. Acid Orange 7 from aqueous solution by UV irradiation in the presence of ZnO nanopowder[END_REF]Sadik 2007a;Sadik 2007b;[START_REF] Behnajady | Design equation with mathematical kinetic modeling for photooxidative degradation of C.I. Acid Orange 7 in an annular continuous-flow photoreactor[END_REF]. where U cell is the average cell voltage (V), I is the current (A), t is the electrolysis time (h) and V is the volume (L).

The electrical energy consumption for Orange II degradation under different reaction systems were summarized in Table 7-3. It can be seen from Table 7-3 that the electrocatalysis processes were more energy saving than the photocatalysis processes.

For example, EE/O value was 273.9 (kWh m -3 order -1 ) for UV/TiO 2 process, while EC/Fe 3 O 4 process only needed 19.65 (kWh m -3 order -1 ) electrical energy consumption.

What's more, the presence of oxidant (H 2 O 2 or PDS) could improve electrocatalysis and photocatalysis, and the associated electrical energy consumption decreased. Table 4 indicated that the EE/O value of EC/Fe 3 O 4 /PDS process was less than one-half of EC/ Fe 3 O 4 process.

Future perspectives

Further studies can be developed as follows:

(1) Finding out the suitable method to detect the intermediate products of artificial sweeteners and then obtain the degradation pathway;

(2) In the sulfate radical-based electro-Fenton-like process, persufate is remained in the solution after the reaction. Moreover, persulfate can be decomposed to SO 4  and H + , resulting the decline of the pH value. So it is important to separate persulfate and SO 4  from the treated solution.

(3) Both sulfate radicals and hydroxyl radicals are contributed to the degradation

  photoelectron spectroscopy (XPS) was applied to investigate the surface properties of Fe 3 O 4 before and after reaction. GCMS analysis was employed to identify the intermediate products and a plausible degradation pathway of Orange II was proposed.The change of acute toxicity during the treatment was investigated by activated sludge inhibition test. The TOC removal efficiency was 30.0% in a 90 min treatment.

2. 1

 1 Artificial sweeteners 2.2 Synthetic dyes 2.3 Advanced oxidation processes (AOPs) 2.4 Iron metal oxides application in AOPs The following three chapters are the presentation of experiments results related to electro-Fenton process. Chapter 3 is entitled -Treatment of aspartame in aqueous solution by electro-Fenton process‖. The removal of ASP in electro-Fenton process was investigated. The effect of Fe 2+ concentration and applied current on the degradation and mineralization of ASP was evaluated. The absolute rate constant for the reaction between ASP and • OH was determined by using the competition kinetic method. The concentration of the formed carboxylic acids and released inorganic ion was monitored by ion-exclusion chromatography and ion chromatography (IC). The toxicity of ASP and/or its reaction byproducts was studied based on the V. fischeri bacteria luminescence inhibition. Chapter 4 is entitled -Treatment of saccharin (SAC) in aqueous solution by electro-Fenton process‖. The removal of artificial sweeteners saccharin (SAC) in aqueous solution by electro-Fenton processes was performed in this study. Different anode materials on the removal and mineralization of SAC were tested. The absolute rate constant of the SAC hydroxylation reaction was determined. The evolution of carboxylic acids was determined by ion-exchange chromatography. The evolution of toxicity of SAC and/or its reaction byproducts based on the V. fischeri bacteria luminescence inhibition was studied. Chapter 5 is entitled -Treatment of sucralose (SUC) in aqueous solution by electro-Fenton process‖. The removal of artificial sweeteners SUC in aqueous solution by electro-Fenton processes was carried out in this study. The effect of Fe 2+ concentration and applied current on the mineralization of SUC was evaluated. The concentration of the formed carboxylic acids and released inorganic ion was monitored by ion-exchange chromatography and ion chromatography. The toxicity of SUC and/or its reaction byproducts was investigated based on the V. fischeri bacteria luminescence inhibition. The following two chapters are the parts devoted to the experiments results of electro-Fenton-like process. Chapter 6 is entitled -Decolorization of Orange II in water by electro/α-FeOOH/PDS process‖. In this section, the removal of Orange II by electro/α-FeOOH/peroxydisulfate process was investigated. Response surface methodology (RSM) based on Box-Behnken statistical experiment design (BBD) was applied to analyze the experimental variables. The positive and negative effects on the decolorization of Orange II were determined. The response surface methodology models were derived based on the results of the pseudo-first-order decolorization rate constant and the response surface plots were developed accordingly. Chapter 7 is entitled -Decolorization of Orange II in water by electro/Fe 3 O 4 /PDS process‖. In this chapter, the decolorization of Orange II was performed in aqueous solution by Fe 3 O 4 activated peroxydisulfate (PDS) oxidation in an electrochemical reactor (EC/Fe 3 O 4 /PDS process) was performed. Various parameters were investigated to optimize the process. The surface properties of Fe 3 O 4 before and after reaction was investigate. The intermediates of Orange II during the treatment were identified and the degradation pathway was proposed. The change of acute toxicity during the treatment was investigated.

  photobleaching of Orange II within seconds by the oxone/Co 2+ reagent in Fenton-like processes. Applied Catalysis B: Environmental 49(3), 207-215. Fernandez, J., Nadtochenko, V. and Kiwi, J. (2003). Photobleaching of Orange II within seconds using the oxone/Co 2+ reagent through Fenton-like chemistry. Chemical Communications 18, 2382-2383. Garcí a, O., Isarain-Chá vez, E., El-Ghenymy, A., Brillas, E. and Peralta-Herná ndez, J.M. (2014). Degradation of 2,4-D herbicide in a recirculation flow plant with a Pt/air-diffusion and a BDD/BDD cell by electrochemical oxidation and electro-Fenton process. Journal of Electroanalytical Chemistry 728, 1-9. Gayathri, P., Praveena Juliya Dorathi, R. and Palanivelu, K. (2010). Sonochemical degradation of textile dyes in aqueous solution using sulphate radicals activated by immobilized cobalt ions. Ultrasonics Sonochemistry 17(3), 566-571. Ghoneim, M.M., El-Desoky, H.S. and Zidan, N.M. (2011). Electro-Fenton oxidation of Sunset Yellow FCF azo-dye in aqueous solutions. Desalination 274(1-3), 22-30. Hermosilla, D., Cortijo, M. and Huang, C.P. (2009a). Optimizing the treatment of landfill leachate by conventional Fenton and photo-Fenton processes. Science of The Total Environment 407(11), 3473-3481. Hermosilla, D., Cortijo, M. and Huang, C.P. (2009b). The role of iron on the degradation and mineralization of organic compounds using conventional Fenton and photo-Fenton processes. Chemical Engineering Journal 155(3), 637-646.

Figure 2 - 1

 21 Figure 2-1 Schematic representation of different categories of EAOPs based on Fenton's reagent ((a) electro-Fenton process; (b) Fered Fenton process; (c) electrochemical peroxidation process;)

Figure 2 - 2

 22 Figure 2-2 Schematic representation of electro-Fenton cells divided by salt bridge (a) and ion exchange membrane (b)

  T). The EFO process demonstrated high 2,4,5-T removal efficiency and nearly completed the herbicide degradation within 10 min. Yuan et al. 2013 employed a similar system to degrade trichloroethylene (TCE).

  -15) and[START_REF] Heng | Degradation of clofibric acid in aqueous solution by an EC/Fe 3+ /PMS process[END_REF][START_REF] Heng | Degradation of bisphenol A in aqueous solution by a novel electro/Fe 3+ /peroxydisulfate process[END_REF][START_REF] Heng | Electro-Fenton removal of Orange II in a divided cell: Reaction mechanism, degradation pathway and toxicity evolution[END_REF][START_REF] Heng | Degradation of Orange II in aqueous solution by a novel electro/Fe 3 O 4 process[END_REF][START_REF] Heng Lin | Treatment of organic pollutants using electro-Fenton and electro-Fenton-like process in aqueous solution[END_REF](7)(8)(9)(10)(11)(12)(13)(14)(15)(16)) can transform to Fe 2+ via cathodic reduction (Eq.(2-12)) (Figure2-3b).Wu et al. 2012 chose azo dye Orange II as the target pollutant and treated it using an -electro/Fe 2+ /peroxydisulfate‖ process. 95% of 0.1 mM Orange II can be removed in a 60 min reaction with electrolyte concentration 0.1 M, Fe 2+ concentration 1 mM, PDS concentration 12 mM, and current density (j) 16.8 mA/cm 2 .

Figure 2 - 3

 23 Figure 2-3 Schematic representation of different types of electro-Fenton-like processes: (a) The first category; (b) The second category; (c) The third category)

Fe 3 O 4

 34 photo-Fenton and TiO 2 photocatalytic degradation of the artificial sweetener sucralose in aqueous media. Applied Catalysis B: Environmental 129, 71-79. Cao, J., Zhang, W.-X., Brown, D.G. and Sethi, D. (2008) Oxidation of lindane with Fe (II)-activated sodium persulfate. Environmental Engineering Science 25(2), 221-228. Chang, J.-S., Chou, C., Lin, Y.-C., Lin, P.-J., Ho, J.-Y. and Lee Hu, T. (2001) Kinetic characteristics of bacterial azo-dye decolorization by Pseudomonas luteola. Water Research 35(12), 2841-2850. Chen, K., Wang, G.-H., Li, W.-B., Wan, D., Hu, Q. and Lu, L.-L. (2014). Application of response surface methodology for optimization of Orange II removal by heterogeneous Fenton-like process using Fe 3 O 4 nanoparticles. Chinese Chemical Letters 25(11), 1455-1460. Chou, S., Huang, Y.-H., Lee, S.-N., Huang, G.-H. and Huang, C. (1999) Treatment of high strength hexamine-containing wastewater by electro-Fenton method. Water Research 33(3), 751-759. Chun, J., Lee, H., Lee, S.-H., Hong, S.-W., Lee, J., Lee, C. and Lee, J. (2012) Magnetite/mesocellular carbon foam as a magnetically recoverable fenton catalyst for removal of phenol and arsenic. Chemosphere 89(10), 1230-1237. Cleveland, V., Bingham, J.-P. and Kan, E. (2014) Heterogeneous Fenton degradation of bisphenol A by carbon nanotube-supported Fe 3 O 4 . Separation and Purification Technology 133, 388-395.

H 2 O

 2 2 was supplied from the electro-reduction of O 2 dissolved in the solution in all the electrolyses. The degradation experiments were performed using 0.2 mM ASP solution with 50 mM Na 2 SO 4 as supporting electrolyte. A catalytic quantity of ferrous ion was added into the solution before the beginning of electrolysis. The initial pH (pH 0 ) of ASP solutions was measured with a CyberScan pH 1500 pH-meter (Eutech Instrument, USA)

  ions (NO 3 -and NH 4 + ) released during the treatment were monitored by ion chromatography with a Dionex ICS-1000 Basic Ion Chromatography System equipped with an IonPac AS4A-SC (anion exchange) and CS12 A (cation exchange) 250 mm × 4 mm column and fitted with a DS6 conductivity detector containing a cell heated at 35 C under control through a Chromeleon SE software. The mobile phase was a mixture of 3.6 mM Na 2 CO 3 and 3.4 mM NaHCO 3 solution for anion-exchange column with a flow rate of 2.0 mL/min and 9 mM H 2 SO 4 solution for cation-exchange with a flow rate of 1.0 mL/min. The volume of injections was 25 μL.

  -1, ASP can be totally removed in a 20 min reaction using Pt anode, while it only needs 15 min to degrade ASP completely with BDD anode. This indicated that the use of BDD did lead to the acceleration of the ASP degradation. On the one hand, due to the low adsorption ability of • OH on BDD, the loosely bound BDD( • OH) formed at the anode surface (Eq. (3-1)) can readily react with organic pollutant, in contrast to the chemisorbed radicals typically formed at Pt surface which limited the oxidation ability of Pt( • OH)(Özcan et al. 2009a; Dirany et al. 2012). On the other hand, reactive BDD( • OH) is generated in much higher quantities than Pt( • OH) when electrolysis is operated at the current within the water discharge region(Brillas et al. 2005; Panizza and Cerisola 2005; Brillas et al. 2009) because of the large O 2 evolution overvoltage of the former. BDD + H 2 O →BDD( • OH) + H + + e  (3-1)

  concentration might be attributed to the role of Fe 2+ as scavenger of hydroxyl radicals (Eq. (3-2)) which takes place with a large rate constant (k = 3.20 × 10 8 M 1 S 1 )(Siré s et al. 2007b; Oturan et al. 2010).

Fig. 3 - 1

 31 Fig. 3-1 Effect of Fe 2+ (as catalyst) concentration on electro-Fenton degradation of 0.2 mM ASP with Pt (a) and BDD (b) anode versus carbon-felt cathode. Experimental conditions: I = 200 mA, pH 0 = 3.0, [Na 2 SO 4 ] = 50 mM

  ). This means the application of BDD anode can improve removal efficiency of ASP and reduce energy consumption by using lower current intensity. A similar behavior has been observed for the degradation of sulfachloropyridazine(Dirany et al. 2012). At current intensity above 200 mA, the removal efficiency decreased. This could due to the increase of side reactions: (i) H 2 evolution from H 2 O reduction (Eq.[START_REF] Heng | Degradation of bisphenol A in aqueous solution by a novel electro/Fe 3+ /peroxydisulfate process[END_REF][START_REF] Heng | Electro-Fenton removal of Orange II in a divided cell: Reaction mechanism, degradation pathway and toxicity evolution[END_REF][START_REF] Heng | Degradation of Orange II in aqueous solution by a novel electro/Fe 3 O 4 process[END_REF]). (ii) The 4-e -reduction of O 2 leading to the formation of H 2 O (Eq.[START_REF] Heng | Degradation of bisphenol A in aqueous solution by a novel electro/Fe 3+ /peroxydisulfate process[END_REF][START_REF] Heng | Electro-Fenton removal of Orange II in a divided cell: Reaction mechanism, degradation pathway and toxicity evolution[END_REF][START_REF] Heng | Degradation of Orange II in aqueous solution by a novel electro/Fe 3 O 4 process[END_REF][START_REF] Heng Lin | Treatment of organic pollutants using electro-Fenton and electro-Fenton-like process in aqueous solution[END_REF]) which inhibited the H 2 O 2 formation reaction (Eq.(3-4))

Fig. 3 - 2

 32 Fig. 3-2 Effect of applied current on electro-Fenton degradation of 0.2 mM ASP by Pt (a) and BDD (b) anode versus carbon-felt cathode. Experimental conditions: [Fe 2+ ] = 0.2 mM, pH 0 = 3.0, [Na 2 SO 4 ] = 50 mM

Fig. 3 - 3 Fig. 3 -

 333 Fig. 3-3 Determination of k abs(ASP) by competition kinetic method according to the pseudo first-order kinetics between ASP and BA during the electro-Fenton process. Experimental conditions: [ASP] 0 = 0.1 mM, ([BA] 0 = 0.1 mM, [Fe 2+ ] = 0.2 mM, I = 50 mA, pH 0 3.0, [Na 2 SO 4 ] = 50 mM

0. 1 Fig. 3 - 4

 134 Fig. 3-4 Effect of Fe 2+ concentration on mineralization of 0.2 mM ASP with Pt (a) and BDD (b) anode versus carbon-felt cathode. Experimental conditions: I = 200 mA, pH 0 3.0, [Na 2 SO 4 ] = 50 mM

Fig. 3 - 5

 35 Fig. 3-5 Effect of applied current on mineralization of 0.2 mM ASP with Pt (a) and BDD (b) anode versus carbon-felt cathode. Experimental conditions: [Fe 2+ ] = 0.2 mM, pH 0 3.0, [Na 2 SO 4 ] = 50 mM

Fig. 3 - 6 3 (Fig. 3 - 7

 36337 Fig. 3-6 Evolution of short-chain carboxylic acids formed during electro-Fenton processes with Pt (a) and BDD (b) anodes versus carbon-felt cathode. Experimental conditions: [ASP] 0 = 0.2 mM, [Fe 2+ ] = 0.2 mM, I = 200 mA, pH 0 3.0, [Na 2 SO 4 ] = 50 mM

Fig. 3 - 8 4 Conclusions

 384 Fig. 3-8 Evolution of the inhibition of marine bacteria, Vibrio fischeri luminescence (Microtox method) during electro-Fenton processes. Experimental conditions: [ASP] 0 = 0.2 mM, [Fe 2+ ] = 0.2 mM, I = 200 mA, pH 0 3.0, [Na 2 SO 4 ] = 50 mM

Fig. 4 - 1

 41 Fig. 4-1 Effect of the nature of supporting electrolyte on the oxidative degradation of SAC with DSA (a), Pt (b) and BDD (c) anodes ([SAC] 0 = 0.2 mM, [Fe 2+ ] = 0.2 mM, I = 200 mA, pH 0 3.0)

Fig. 4 - 2

 42 Fig. 4-2 Comparison of the performance of different anode in the degradation (a) and mineralization (b) of SAC ([SAC] 0 = 0.2 mM, [Fe 2+ ] = 0.2 mM, [Na 2 SO 4 ] = 50 mM, I = 200 mA, pH 0 3.0)

FeFig. 4 - 3

 43 Fig. 4-3 Effect of Fe 2+ (as catalyst) concentration on the removal of SAC with Pt (a) and BDD (b) anode ([SAC] 0 = 0.2 mM, [Na 2 SO 4 ] = 50 mM, I = 200 mA, pH 0 3.0)

Fig. 4 - 4

 44 Fig. 4-4 Effect of applied current on the destruction kinetics of SAC with Pt (a) and BDD (b) anode ([SAC] 0 = 0.2 mM, [Fe 2+ ] = 0.2 mM, [Na 2 SO 4 ] = 50 mM, pH 0 3.0)

Fig. 4 - 5 4 . 3 . 6

 45436 Fig. 4-5 Determination of the absolute constant of reaction between SAC and • OH using competiton kinetics method by taking BA as standard competitors ([SAC] 0 = 0.1 mM, ([BA] 0 = 0.1 mM, [Fe 2+ ] = 0.2 mM, I = 50 mA, pH 0 3.0, [Na 2 SO 4 ] = 50 mM)

Figure 4 -Fig. 4 - 6

 446  the optimal Fe 2+ concentration for both Pt/carbon-felt and BDD/carbon-felt cells was 0.2 mM. When the Fe 2+ concentration was higher than 0.2 mM, the percentage of scavenged hydroxyl radicals from Fe 2+ ions increased, thus the side reactions between Fe 2+ ions and • OH (Eq. (4-7)) became more significant(Özcan et al. 2008a).It can be seen by comparing Fig.4-6a and Fig.4-6b that the mineralization of SAC was accelerated by the employ of BDD anode. At optimal conditions, BDD/carbon-felt cell achieved 93.3% mineralization in only 180 min, while this rate was about 70% for or Pt/carbon-felt cell at same time. A mineralization rate of 76.1% was reached by this cell at 360 min. BDD anode has the much higher O 2 evolution overpotential(1.27 V) than Pt (0.27 V)(Oturan et al. 2011). So the oxidative action of BDD( • OH) is much more efficient than Pt( • OH) (Martí nez-Huitle and Brillas 2009). Moreover, the loosely bound BDD( • OH) formed at the anode surface (Eq.[START_REF] Heng | Electro-Fenton removal of Orange II in a divided cell: Reaction mechanism, degradation pathway and toxicity evolution[END_REF][START_REF] Heng | Degradation of Orange II in aqueous solution by a novel electro/Fe 3 O 4 process[END_REF][START_REF] Heng Lin | Treatment of organic pollutants using electro-Fenton and electro-Fenton-like process in aqueous solution[END_REF]) can readily react with organic pollutant due to the low adsorption ability of • OH on BDD, in contrast to the chemisorbed radicals Pt( • OH) which relatively strongly attached to the surface and less reactive(Oturan et al. 2011; Dirany et al. 2012).

Fig. 4 - 7

 47 Fig. 4-7 Effect of applied current on mineralization of SAC by electro-Fenton process with Pt (a) and BDD (b) anode ([SAC] 0 = 0.2 mM, [Fe 2+ ] = 0.2 mM, [Na 2 SO 4 ] = 50 mM, pH 0 3.0)

Fig. 4 - 8

 48 Fig. 4-8 Evolution of short-chain carboxylic acids formed during electro-Fenton processes using Pt (a) and BDD (b) anodes ([SAC] 0 = 0.2 mM, [Fe 2+ ] = 0.2 mM, , [Na 2 SO 4 ] = 50 mM, I = 200 mA, pH 0 3.0)

  0.2 mM SAC solution was examined at 200 mA constant current.The evolution of the luminescence inhibition as function of the electrolysis time for exposition time of 5 and 15 min are tested. Only the curves obtained at both Pt and BDD anodes after a 15-min exposure time were presented in Fig.4-9, because the curves recorded after a 5-min exposure time were very similar. For both Pt and BDD anode, the toxicity increased significantly and reached the maximum luminescence inhibition peak at 30 min, indicating the formation of significantly more toxic by-products at the beginning of the treatment. The inhibition ratio decreased after 30 min, in relation to the destruction of aromatics and the increase of the less toxic by-products. These less toxic byproducts mainly included the carboxylic acids, which present a very low toxicity toward V. fischeri bacteria(Dirany et al. 2012).

Fig. 4 - 9

 49 Fig. 4-9 Inhibition of luminescence of V. fischeri bacteria during electro-Fenton processes with Pt and BDD anode ([SAC] 0 = 0.2 mM, [Fe 2+ ] = 0.2 mM, I = 200 mA, pH 0 3.0, [Na 2 SO 4 ] = 50 mM)

  ). Reaction kinetics and efficiencies for the hydroxyl and sulfate radical based oxidation of artificial sweeteners in water. The Journal of Physical Chemistry A 116(40), 9819-9824. Yahya, M.S., Oturan, N., El Kacemi, K., El Karbane, M., Aravindakumar, C.T. and Oturan, M.A. (2014) Oxidative degradation study on antimicrobial agent ciprofloxacin by electro-fenton process: Kinetics and oxidation products.

Fig. 5 - 1

 51 Fig. 5-1 Effect of Fe 2+ concentration on mineralization of SAC during electro-Fenton process using Pt (a) and BDD (b) anode ([SUC] 0 = 0.2 mM, [Na 2 SO 4 ] = 50 mM, I = 200 mA, pH 0 3.0)

Fig. 5 -Fig. 5 - 2

 552 Fig. 5-2 Effect of applied current on mineralization of SAC by electro-Fenton process with Pt (a) and BDD (b) anode ([SUC] 0 = 0.2 mM, [Fe 2+ ] = 0.2 mM, [Na 2 SO 4 ] = 50 mM, pH 0 3.0)

Fig. 5 - 3

 53 Fig. 5-3 Dependence of mineralization current efficiency calculated from Eq. (5-10) on the electrolysis time for the experiments reported in: (a) Fig. 5-2a and (b) Fig. 5-2b

Fig. 5 - 4

 54 Fig. 5-4 Evolution of short-chain carboxylic acids formed during electro-Fenton processes with Pt v(a) and BDD (b) anodes ([SUC] 0 = 0.2 mM, [Fe 2+ ] = 0.2 mM, [Na 2 SO 4 ] = 50 mM, I = 200 mA, pH 0 3.0)

Fig. 5 - 6

 56 Fig. 5-6 Evolution of the inhibition of marine bacteria, Vibrio fischeri luminescence (Microtox method) during electro-Fenton processes with Pt (a) and BDD (b) anode ([sucralose] 0 = 0.2 mM, [Fe 2+ ] = 0.2 mM, [Na 2 SO 4 ] = 50 mM, I = 200 mA, pH 0 3.0)

1

 1 Decolorization of Orange II under different systemsIn order or investigate the decolorization efficiency of Orange II under different oxidation systems, 50 mg/L Orange II solution was treated for 120 min by PDS alone, α-FeOOH alone, α-FeOOH/PDS process, EC alone, EC/α-FeOOH process, EC/PDS process and EC/α-FeOOH/PDS process. The results were shown in Fig.6-1. It can be seen from Fig.6-1a that little decolorization occurred when Orange II was treated by PDS alone, because PDS is stable at ambient temperature and its oxidation power is limited (E 0 = 2.01 V)( Zhao et al. 2010;[START_REF] Wu | Degradation of Acid Orange 7 in aqueous solution by a novel electro/Fe 2+ /peroxydisulfate process[END_REF], which could hardly degrade Orange II. What' more, the remaining percentage of PDS after 120 min reaction in the PDS alone process was 98.2% (Fig.6-1b), indicating almost no PDS decomposed in this case. Negligible color removed was also observed when α-FeOOH alone was applied, indicating the effect of adsorption of Orange II decolorization was not obvious under the condition investigated. Orange II was hardly removed in α-FeOOH/PDS process and the remaining percentage of PDS was 95.7%, beause PDS can not be activated by Fe(III) on the surface of α-FeOOH. When treated by EC alone, 51.8% of 50 mg/L Orange II was decolorizated because a certain amount of hydroxyl radicals (M( • OH)) was formed as intermediate of water discharge on DSA anode (M) (Eq.

Fig. 6 - 2

 62 Fig. 6-2 XRD pattern of α-FeOOH (Reference X-ray lines belonging to α-FeOOH (JCPDS card n° 29-0713) are given at the bottom of the figure) XPS was used to applied to observe the surface properties of α-FeOOH before and after reaction. The results were showed in Fig. 6-3. The domain peak for O 1s was at 531.6 eV. The O 1s peak of α-FeOOH is shifted significantly higher (0.4 eV) than that of α-Fe 2 O 3 (531.2 eV)(Guo et al. 2010), indicating the proton on the hydroxyl oxygen has some interaction with the O 2 -oxygen (McIntyre and Zetaruk 1977). Fig. 6-3a showed the O 1s peak of α-FeOOH did not change obviously after the reaction. So the properties of oxygen on the surface of α-FeOOH did not change during the EC/α-FeOOH/PDS process. The peaks of Fe 2p spectrum at 711, 719, and 725 eV represented the binding energies of Fe 2p3/2, shake-upsatellite Fe 2p3/2, and Fe 2p1/2, respectively (Nie et al. 2009). As can be seen from Fig. 6-3b, there was no Fe (II) on the surface of α-FeOOH before the reaction. However, The ratio of Fe(II)/Fe(III) on the surface of α-FeOOH adsorbent increased to 0.80 after 120 min reaction. This implied more Fe(II) was generated on the surface of α-FeOOH particles during EC/α-FeOOH/PDS process (Eq. (6-5)).

Fig. 6 - 3

 63 Fig. 6-3 O 1s (a) and Fe2p (b) XPS spectra of fresh and used α-FeOOH after the reaction

Fig. 6 - 4

 64 Fig. 6-4 Proposed mechanism of EC/α-FeOOH/PDS process

Fig. 6 - 5

 65 Fig. 6-5 The effect of initial pH value on the degradation of Orange II (C 0 = 50 mg/L, [α-FeOOH] = 0.5 g/L, [PDS] = 2 g/L, I = 0.5 A, [Na 2 SO 4 ] = 50 mM)

Fig. 6 - 6

 66 Fig. 6-6 Main effects plot for Orange II decolorization rate constant: Applied current (A); α-FeOOH dosage (B); PDS concentration (C)

Fig. 6 - 7

 67 Fig. 6-7 Interaction plots for Orange II decolorization rate constant: Applied current (A); α-FeOOH dosage (B); PDS concentration (C)

Fig. 6 - 8 Fig. 6 - 9 Fig. 6 -

 68696 Fig. 6-8 The internally studentized residuals and normal % probability plot for Orange II decolorization rate constant

Fig. 6 -

 6 Fig. 6-11 Response surface showing removal efficiency as a function of two independent variables: (a) electrical current (A) and α-FeOOH dosage (B), (b) electrical current (A) and PDS concentration (C), (c) α-FeOOH dosage (B) and and PDS concentration (C)

Fig. 6 - 6 . 3 . 5

 6635 Fig. 6-12 Recycling study of α-FeOOH during Orange II degradation (C 0 = 50 mg/L, [α-FeOOH] = 0.5 g/L, [PDS] = 2 g/L, I = 0.5 A, pH 0 7, [Na 2 SO 4 ] = 50 mM)

Fig. 6 -

 6 Fig. 6-13 UV-visible spectra changes with reaction time in EC/α-FeOOH/PDS process (a) and the changes of TOC removal efficiency with reaction time (b) (C 0 = 50 mg/L, [α-FeOOH] = 0.5 g/L, I = 0.5 A, pH 0 7, [Na 2 SO 4 ] = 50 mM)

  keep the temperature constant at 25 C. A Model RYI-3010 direct current (DC) power supply from Shenzhen Yizhan Source, Co., Ltd. (China) was employed to provide the constant current. Before each run, a fresh stock solution of Orange II was prepared in deionised water and the initial concentration (C 0 ) was kept at 25 mg/L. Sodium sulfate (50 mM) was added as a background electrolyte. The initial pH (pH 0 ) of Orange II solutions was measured with a Mettler-Toledo FE20 pH-meter (Mettler-Toledo Instruments Co., Ltd., Shanghai) and adjusted by sulfuric acid or sodium hydroxide. Before the beginning of electrolysis, a certain amount of PDS and Fe 3 O 4 were added into 200 mL of Orange II solution with constant stirring by using a mechanical stirrer (RW 20, IKA, Germany).

Fig. 7 - 1

 71 Fig. 7-1 Schematic of the experimental setup employed in the present study

  Fig. 7-3 O 1s (a) and Fe2p (b) XPS spectra of fresh and used Fe 3 O 4 after the reaction

Fig. 7 - 4

 74 Fig. 7-4 Proposed mechanism of EC/ Fe 3 O 4 /PDS process

  pH of Orange II solution after 60 min at different initial pH was almost the same (about 2.3). As a result, the decolorization efficiency of Orange II and the decomposition of PDS (Table7-1) were similar at different initial pH values. This proved that the EC/Fe 3 O 4 /PDS process could be successfully applied at a wide range of initial pH from 3.0 to 9.0. Since pH 0 6.0 was the natural pH of 25 mg/L Orange II solution, it was selected in the following experiments.

Fig. 7 - 5 7. 3 . 3

 7533 Fig. 7-5 The effect of initial pH on the oxidation efficiency of Orange II (C 0 = 25 mg/L, [PDS] = 20 mM, [Fe 3 O 4 ] = 0.5 g/L, j = 8.4 mA/cm 2 , [Na 2 SO 4 ] = 50 mM)

Fig. 7 - 6

 76 Fig. 7-6 The effect of current density on the oxidation efficiency of Orange II (C 0 = 25 mg/L, [PDS] = 20 mM, [Fe 3 O 4 ] = 0.5 g/L, pH 0 6.0, [Na 2 SO 4 ] = 50 mM)

Fig. 7 - 8 Fig. 7 - 9

 7879 Fig. 7-8 The effect of Fe 3 O 4 dosage on the oxidation efficiency of Orange II (C 0 = 25 mg/L, [PDS] = 10 mM, pH 0 6.0, j = 8.4 mA/cm 2 , [Na 2 SO 4 ] = 50 mM)

Fig. 7 -

 7 Fig. 7-10 Proposed pathway for the oxidation of Orange II by EC/Fe 3 O 4 /PDS process

Fig. 7 -

 7 Fig. 7-11 The changes of TOC, Orange II concentration and SOUR with reaction time (C 0 = 25 mg/L, [PDS] = 10 mM, [Fe 3 O 4 ] = 0.8 g/L, j = 16.8 mA/cm 2 , pH 0 6.0, [Na 2 SO 4 ] = 50 mM)

EE

  2 /H 2 O 2 [Orange II] 0 = 17.5 mg/L, [H 2 O 2 ] = 71 mM, [TiO 2 ] = 0.5 g/L, , pH 0 6.1, 43 W UV lamp 40.45 Sadik 2007b UV/TiO 2 +ZnO/PDS [Orange II] 0 = 17.5 mg/L, [PDS] = 8.5 mM, [TiO 2 +ZnO] 18.08 Sadik = 0.25 g/L, , pH 0 6.1, 43 W UV lamp 2007a EC/Fe 3 O 4 /PDS [Orange II] 0 = 25 mg/L, [PDS] = 10 mM, [Fe 3 Aqueous solutions of Orange II have been degraded effectively in the EC/Fe 3 O 4 /PDS process, in which Fe 3 O 4 -activated PDS process is enhanced by the electrochemical process. The effect of initial pH, current density, PDS concentration and Fe 3 O 4 dosage, on the Orange II decolorization was investigated. The results indicated that Orange II can be totally decolorizated in a 60 min reaction when initial Orange II concentration was 25 mg/L, PDS concentration was 10 mM, Fe 3 O 4 dosage was 0.8 g/L, current density was 8.4 mA/cm 2 and initial pH was 6.0. Recycle experiments showed Fe 3 O 4 particles were stable and can be reused. XPS spectrum indicated Fe(II) was generated on the surface of Fe 3 O 4 particles after reaction. The main intermediates were separated and identified by GCMS technique and a plausible degradation pathway of Orange II was proposed. based on Box-Behnken statistical experiment design was applied to analyze the experimental variables. The response surface methodology models were derived based on the results of the pseudo-first-order decolorization rate kinetics and the response surface plots were developed accordingly. The results indicated that the applied current showed a positive effect on the decolorization rate constant of Orange II. The interaction of α-FeOOH dosage and PDS concentration were significant. The ANOVA results confirmed that the proposed models were accurate and reliable for the analysis of the varibles of EC/α-FeOOH/PDS process. The catalyst α-FeOOH showed good structural stability and could be reused. (5) Aqueous solutions of Orange II have been degraded effectively in the EC/Fe 3 O 4 /PDS process. The decolorization rate was affected by the initial pH of Orange II solution, current density, PDS concentration and Fe 3 O 4 dosage. Orange II can be totally decolorized in a 60 min reaction time when initial Orange II concentration was 25 mg/L, PDS concentration was 10 mM, Fe 3 O 4 dosage was 0.8 g/L, current density was 8.4 mA/cm 2 and initial pH was 6.0. Recycle experiments showed that Fe 3 O 4 particles were stable and can be reused. XPS spectrum indicated that Fe(II) was generated on the surface of Fe 3 O 4 particles after reaction. The main reaction intermediates of Orange II were separated and identified by GCMS technique and a plausible degradation pathway was proposed.

Table 2 - 1

 21 The chemical structure and main characteristics of some commonly used artificial sweeteners

						Molecular weight			Number of times
	Name	Chemical structure	Molecular formula		CAS number	Water solubility (g/L)
						(g/mol)			sweeter than sucrose
	Aspartame	O	O	OCH3	C 14 H 18 N 2 O 5	394.31	22839-47-0	~10 (25 C)	200
			N H
		OH	NH 2	O
			O
	Saccharin			NH	C 7 H 5 NO 3 S	183.18	81-07-2	4	200-700
			S
			O	O
		HO
				OH
		Cl	HO
	Sucralose		O	Cl	C 12 H 19 Cl 3 O 8	397.63	56038-13-2	283 (20 C)	600
		HO	O	O
			OH	Cl
			O
	Acesulfame		HN		C 4 H 5 NO 4 S	163.15	33665-90-6	270 (20 C)	200
		O	S
			O	CH 3
			O

Advanced Oxidation Processes (EAOPs) based on Fenton's reaction

	eco-friendly methods and avoid some disadvantages in conventional Fenton process that heterogeneous hydroxyl radicals M( • OH) electro-generated on high oxygen overvoltage
	have recently attracted great attention for water remediation (Nidheesh and anodes (Eq. (2-13)).
	Gandhimathi 2012). EAOPs based on Fenton's reaction chemistry can be generally M (H 2 O) → M ( • OH) + H + + e  (2-13)
	divided into three categories. Dirany et al. 2012 treated antibiotics sulfachloropyridazine (SCP) in aqueous
	The first one is electro-Fenton process. H 2 O 2 is formed by the reduction of the solution using electro-Fenton process. When the applied current was 300 mA, the initial
	dissolved oxygen on the cathode surface (Eq. (2-11)) in an electrolytic cell. H 2 O 2 can pH of 0.21 mM SCP solution was fixed at 3.0, SCP could be totally removed in 10 min.
	then react with the externally added catalyst (Fe 2+ or Fe 3+ ) to produce • OH (Eq. (2-1)). The mineralization efficiency of SCP was nearly 100% in a 360 min reaction (Table
	Moreover, Fe 2+ which consumed in Fenton's reaction can be regenerated by cathodic 2-2).
	reduction (Eq. (2-12)) in electro-Fenton process which reduces the required The second category is Fered-Fenton process or EF-Fere process. Both
	concentration of initial Fe 2+ to an catalytic amount (Figure 2-1a). components of Fenton's reagent (H 2 O 2 and Fe 2+ ) are added externally to the reactor. Fe 3+
	O 2 + 2H + + 2e  → H 2 O 2 which formed in the Fenton's reaction can convert to Fe 2+ via cathodic reduction (Eq. (2-11)
	Fe 3+ + e  → Fe 2+ (2-12)) (Figure 2-1b). Zhang et al. 2007 degraded 4-nitrophenol (4-NP) in water by (2-12)
	Electro-Fenton process solves some problems existed in Fenton process. For example, using Fered-Fenton process. The COD removal efficiency for 200 mg/L 4-NP was 70%
	higher degradation rate of organic pollutants is obtained because of the continuous in a 120 min reaction, while the COD removal efficiency was only 50% in Fenton
	regeneration of Fe 2+ at the cathode and the on-site production of H 2 O 2 avoids its process at the same concentration of Fenton's reagent (H 2 O 2 and Fe 2+ ) (Table 2-2).
	dangerous transport and storage (Martí nez-Huitle and Brillas 2009) and avoiding also The third category is electrochemical peroxidation process. A sacrificial iron anode
	the competition of wasting reactions (2-6), (2-7) and (2-10) (Oturan and Aaron 2014, was used for Fe 2+ electrogeneration from anodic dissolution via Eq. (2-14). H 2 O 2 is
	Siré s et al. 2014).
	The reporting cathode materials favored the electrogeneration of H 2 O 2 were gas
	2.3.2.1 Fundamentals and categories diffusion electrodes (GDEs) (Brillas et al. 1998; Panizza and Cerisola 2009; Barros et al.
	2014; Yu et al. 2015), graphite (Yuan and Lu 2005; Yuan et al. 2006) and
	The electrochemical (EC) technology which based on the transfer of electrons, three-dimensional electrodes such as carbon-felt (Pimentel et al. 2008; Diagne et al.
	have received great attention for the prevention of pollution problems (Brillas et al. 2014; Olvera-Vargas et al. 2014), activated carbon fiber (ACF) (Wang et al. 2005; Lei et
	2009; Oturan and Aaron 2014). The main advantage of EC technology is its al. 2010; Wang et al. 2010a), reticulated vitreous carbon (RVC) (Xie and Li 2006;
	environmental compatibility since the main reagent, electron, is a clean reagent. A large Martí nez and Bahena 2009) and carbon sponge (Özcan et al. 2008). Among these
	variety of EAOPs were developed by combining EC technology with AOPs in the last cathode materials, carbon-felt has a high specific surface that favors the fast generation
	decade due to their environmental safety and compatibility (operating at mild of Fenton's reagent (H 2 O 2 and Fe 2+ ) and has applied widely (Brillas et al. 2009). The
	conditions), versatility, high efficiency, and possibility of automation (Oturan and commonly used anode materials in electro-Fenton process were high oxygen
	Aaron 2014). Among all the EAOPs, EAOPs based on Fenton's reaction chemistry are overvoltage anodes (M), such as PbO 2, dimensionally stable anodes (DSA), Pt and
	boron-doped diamond (BDD) anode. Organic pollutants can also be destroyed by

Table 2

 2

		-2 Examples of different categories of EAOPs based on Fenton's reagent
	Categories	Compound	Typical condition	Corresponding results Ref.
	Electro-Fenton	SCP	BDD anode,CF cathode,[SCP] 0 =	100% removal in 10	Dirany
			0.21mM,[Fe 2+ ] = 0.2mM,I = 300	min and near 100 %	et al.
			mA, pH 0 3.0	TOC removal in 360	2012
				min
	Fered-Fenton	4-NP	Stainless steel cylinder cathode,	About 70% COD	Zhang
			RuO 2 /IrO 2 anode,[4-NP] 0 = 200	removal in a 120 min	et al.
			mg/L,[H 2 O 2 ] = 9.12 mM,H 2 O 2	reaction	2007
			feeding time 60 min,Fe 2+ /H 2 O 2 =
			0.05,I = 1.0 A,pH 0 5.0
	electrochemical	HCB	Stainless steel electrode,gap	96.96% HCB removall	Xie
	peroxidation		between electrodes 6 cm,30%	in a 3 h reaction	et al.
			H 2 O 2 concentration 500 mg/L,pH 0		2005
			2.5
	2.3.2.2 Reactors for EAOPs based Fenton's reaction
	Generally, the reactors in which EAOPs based Fenton's reaction were carried out

Table 3 -

 3

1 Apparent rate constants (k app ) obtained in electro-Fenton processes for ASP degradation, assuming pseudo-first order kinetic model under different operating conditions. Operating conditions: I and [Fe 2+ ] variable, pH: 3.0 and carbon felt

cathode Anode [Fe 2+ ]/mM I/mA k app (min -1 ) R 2

	BDD	0.05	200	0.30±0.02	0.979
	BDD	0.1	200	0.34±0.01	0.993
	BDD	0.2	200	0.51±0.05	0.965
	BDD	0.3	200	0.32±0.01	0.998
	BDD	0.5	200	0.28±0.02	0.989
	BDD	0.2	50	0.15±0.01	0.998
	BDD	0.2	100	0.25±0.01	0.989
	BDD	0.2	300	0.40±0.01	0.998
	BDD	0.2	500	0.25±0.01	0.991
	Pt	0.05	200	0.15±0.01	0.994
	Pt	0.1	200	0.24±0.01	0.992
	Pt	0.2	200	0.34±0.01	0.995
	Pt	0.3	200	0.26±0.01	0.996
	Pt	0.5	200	0.22±0.02	0.999
	Pt	0.2	50	0.14±0.01	0.997
	Pt	0.2	100	0.15±0.01	0.998
	Pt	0.2	300	0.34±0.03	0.968
	Pt	0.2	500	0.17±0.01	0.997

Table 4 - 1

 41 Apparent rate constants (k app ) obtained in electro-Fenton processes for SAC degradation, assuming pseudo-first order kinetic model under different operating conditions.

	Electrode	[Fe 2+ ]/mM	I/mA supporting electrolyte	k app (min -1 )	R 2
	DSA	0.2	200	Na 2 SO 4 (50 mM)	0.181±0.004	0.997
	DSA	0.2	200	NaNO 3 (100 mM)	0.170±0.030	0.998
	DSA	0.2	200	NaCl (100 mM)	0.019±0.001	0.994
	Pt	0.2	200	Na 2 SO 4 (50 mM)	0.211±0.008	0.990
	Pt	0.2	200	NaNO 3 (100 mM)	0.125±0.007	0.977
	Pt	0.2	200	NaCl (100 mM)	0.038±0.001	0.999
	BDD	0.2	200	Na 2 SO 4 (50 mM)	0.194±0.008	0.988
	BDD	0.2	200	NaNO 3 (100 mM)	0.143±0.002	0.998
	BDD	0.2	200	NaCl (100 mM)	0.023±0.001	0.999
	Pt	0.05	200	Na 2 SO 4 (50 mM)	0.047±0.004	0.945
	Pt	0.1	200	Na 2 SO 4 (50 mM)	0.108±0.003	0.995
	Pt	0.3	200	Na 2 SO 4 (50 mM)	0.147±0.012	0.953
	Pt	0.5	200	Na 2 SO 4 (50 mM)	0.063±0.004	0.972
	BDD	0.05	200	Na 2 SO 4 (50 mM)	0.077±0.002	0.994
	BDD	0.1	200	Na 2 SO 4 (50 mM)	0.106±0.003	0.995
	BDD	0.3	200	Na 2 SO 4 (50 mM)	0.117±0.001	0.999
	BDD	0.5	200	Na 2 SO 4 (50 mM)	0.095±0.002	0.996
	Pt	0.2	50	Na 2 SO 4 (50 mM)	0.070±0.001	0.999
	Pt	0.2	100	Na 2 SO 4 (50 mM)	0.176±0.005	0.994
	Pt	0.2	300	Na 2 SO 4 (50 mM)	0.256±0.005	0.998
	Pt	0.2	500	Na 2 SO 4 (50 mM)	0.259±0.005	0.999
	BDD	0.2	50	Na 2 SO 4 (50 mM)	0.089±0.001	0.998
	BDD	0.2	100	Na 2 SO 4 (50 mM)	0.137±0.005	0.991
	BDD	0.2	300	Na 2 SO 4 (50 mM)	0.187±0.006	0.993
	BDD	0.2	500	Na 2 SO 4 (50 mM)	0.156±0.002	0.998

  deoxy-α-Dgalactopyranoside, C 12 H 19 Cl 3 O 8 ) was purchased from Sigma-Aldrich. The chemical structure and main characteristics of SUC were presented in Table 2-1. Analytical grade anhydrous sodium sulfate (supporting electrolyte) and ferrous sulfate heptahydrated (catalyst source) were obtained from Sigma-Aldrich and Acros Organics, respectively.

Table 6 - 1

 61 Chemical structure and main characteristics of Orange II

	Name	Orange II
	Chemical structure
		NaO 3 S	N N	NaO 3 S	H N N
			HO		O
	Molecular formula	C 16 H 11 N 2 NaO 4 S
	CAS number	633-96-5
	Molecular weight	350.32
	λ max (nm)	485
	Water solubility (g/L)	116
	6.2.2 Electrochemical apparatus and procedures
	Batch experiments were conducted in an undivided electrolytic reactor (glass
	beaker) containing 400 mL of Orange II solution. Electrolyses were performed at
	constant current controlled by a direct current (DC) power supply (Model WYK-305)
	from Yangzhou Jintong Source, Co., Ltd. (China). One 5 × 11.9 cm plate anode
	(Ti/RuO 2 -IrO 2 ) and one plate cathode (stainless steel) of the same dimensions were
	arranged parallel to each other at a distance of 4.0 cm. Prior to the electrolysis, the
	initial pH (pH 0 ) of Orange II solutions were measured with a Mettler-Toledo FE20
	pH-meter (Mettler-Toledo Instruments Co., Ltd., Shanghai) and adjusted by sulfuric
	acid and sodium hydroxide. A magnetic stirrer (model 78-1, Hangzhou Instrument

Table 6

 6

	-2. The

Table 6 - 2

 62 Experimental range and levels of the independent variables

	Variables	Symbol	-1	0	+1
	Electric current (A)	X 1	0.2	0.5	0.8
	α-FeOOH dosage (g/L)	X 2	0.2	0.5	0.8
	PDS concentration (g/L)	X 3	1	2	3

Table 6 - 3

 63 Design matrix in coded units and the experiments

						k (min -1 )
	Stand no.	X 1	X 2	X 3	Observed	Predicted
	1	-1	-1	0	0.0079	0.0093
	2	1	-1	0	0.021	0.020
	3	-1	1	0	0.011	0.011
	4	1	1	0	0.022	0.020
	5	-1	0	-1	0.011	0.0092
	6	1	0	-1	0.017	0.017
	7	-1	0	1	0.011	0.011
	8	1	0	1	0.021	0.023

  benzenesulfonic acid sodium salt) with reagent purity grade was obtained from Sinopharm Chemical Reagent Co., Ltd. (China) and used without further purification. Fe 3 O 4 used in this study was from Prolabo Co.

	(Paris, France). Sodium peroxydisulfate (Na 2 S 2 O 8 , 98%) and anhydrous sodium sulfate
	(Na 2 SO 4 , 99%) were purchased from Sinopharm Chemical Reagent Co., Ltd. (Shanghai,
	China) and of analytical grade.

  evaluate the decolorization efficiency of Orange II under different systems, the following blank experiments were conducted: Fe 3 O 4 alone, PDS alone, Fe 3 O 4 /PDS process, EC alone, EC/Fe 3 O 4 process, EC/PDS process and EC/Fe 3 O 4 /PDS process. The results are shown in Fig. 7-2a. It can be obviously seen that no appreciable Orange II decolorization was observed when Fe 3 O 4 was added into the Orange II solution alone,indicating the effect of adsorption on Orange II decolorization was not obvious under the condition investigated. Little decolorization also occurred when PDS alone was applied. PDS is stable at ambient temperature and its oxidation power was limited (E 0 = 2.01 V)[START_REF] Zhao | Enhanced oxidation of 4-chlorophenol using sulfate radicals generated from zero-valent iron and peroxydisulfate at ambient temperature[END_REF][START_REF] Wu | Degradation of Acid Orange 7 in aqueous solution by a novel electro/Fe 2+ /peroxydisulfate process[END_REF], which could hardly degrade Orange II.The decolorization efficiency of Orange II was 93.0% and the remaining percentage of PDS was 33.7% in the process of EC/Fe 3 O 4 /PDS, indicating Fe(II) can be To verify the enhancement of the conversion of Fe(III) into Fe(II) in the process of EC/Fe 3 O 4 /PDS, the X-ray photoelectron spectroscopy was applied to characterize Fe 3 O 4 before and after reaction. The domain peak for O 1s at 530.0 eV is characteristic of oxygen in a metal oxide[START_REF] Zhao | Electro-Fenton oxidation of pesticides with a novel Fe 3 O 4 @Fe 2 O 3 /activated carbon aerogel cathode: High activity, wide pH range and catalytic mechanism[END_REF]. The peak of Fe 2p spectrum at 710.6 and 724.4 eV represent Fe 2p3/2 and 2p1/2, respectively, indicated the existence of Fe 3 O 4[START_REF] Huang | Heterogeneous sono-Fenton catalytic degradation of bisphenol A by Fe 3 O 4 magnetic nanoparticles under neutral condition[END_REF]. As can be seen from Fig.7-3(a), no significantly change was observed about O 1s XPS spectra before and after reaction. Before the reaction, the Fe2p 3/2 spectrum fitted to four peaks located at binding energies (BE): 710.2, 711.3, 712.3 and 714.1 eV (Guo et al. 2010;[START_REF] Zhao | Electro-Fenton oxidation of pesticides with a novel Fe 3 O 4 @Fe 2 O 3 /activated carbon aerogel cathode: High activity, wide pH range and catalytic mechanism[END_REF]. The peak at 710.2 eV could be assigned to the formation of ferrous oxide, where other peaks would be considered as the Fe(III) oxide (Fig.7-3(b)). The ratio of Fe(II)/Fe(III) on the surface of Fe 3 O 4 adsorbent increased from 0.75 to 1.09 after 60 min reaction. This implied more Fe(II) was generated on the surface of Fe 3 O 4 particles in the electrochemical reactor (Eq.

	The remaining percentage of PDS after 60 min reaction in the PDS alone process was
	98.0% (Fig. 7-2b), indicating almost no PDS decomposed in this case. The
	concentration of Orange II was hardly decreased in Fe 3 O 4 /PDS process and the
	remaining percentage of PDS was 90%. This may due to the lack of Fe(II) on the
	surface of Fe 3 O 4 particles. When treated by EC alone, 61.1% of 25 mg/L Orange II was
	decolorizated due to a certain amount of hydroxyl radicals (M( • OH)) formed as
	intermediate of water discharge on DSA anode (M) (Eq. (7-4)) (Özcan et al. 2009).
		M(H 2 O) → M( • OH) + H + + e -	(7-4)
	When 0.5 g/L Fe 3 O 4 was employed to the electrolytic reactor, the decolorization
	efficiency of Orange II increased to 82.8%. It can be explained that in EC/Fe 3 O 4 process,
	Fe 3 O 4 was used as particle electrodes and could increase mass transfer coefficient and
	reduce energy consumption (Lin et al 2013). The decolorization efficiency slightly
	improved to 85.1% in EC/PDS process, which might be attributed to sulfate radicals
	produced via Eq. (7-5). 54.4% of PDS was decomposed in this case (Fig. 7-2b).
	S 2 O 8	2-+ e -→ SO 4	•-+ SO 4	2-	(7-5)

(7-6)).

Table 7

 7

	EC, EC/Fe 3 O 4 , EC/PDS and EC/Fe 3 O 4 /PDS process followed apparent first order
	kinetics and the apparent rate constants (k app ) were 0.015, 0.028, 0.030 and 0.043 min -1 ,
	respectively, at pH 3.0.

-1. It can be seen that the decolorization of 25 mg/L Orange II by

Table 7 - 1

 71 Kinetic coefficients and residual concentrations of PDS under various operating

					conditions
			Current					Residual
	Run pH 0	density (mA/cm 2 )	[PDS] (mM)	[Fe 3 O 4 ] (g/L)	k app (min -1 )	R 2	[PDS] (mM)
	1	3.0	8.40	0	0	0.015	0.998	
	2	3.0	8.40	0	0.5	0.028	0.996	
	3	3.0	8.40	20	0	0.030	0.993	9.12
	4	3.0	8.40	20	0.5	0.043	0.992	6.74

-e - • OH +e - S 2 O 8 2 - H 2 O SO 4 •- OH - Degradation products Fe(II) Fe(III)

	5	6.0	8.40	20	0.5		0.039	0.995	7.56
	6	9.0	8.40	20	0.5		0.037	0.996	7.14
	7	6.0	3.36	20	0.5		0.027	0.988	11.60
	8	6.0	16.8	20	0.5		0.071	0.989	4.04
	9	6.0	25.2	20	0.5		0.117	0.973	2.92
	10	6.0	8.40	5	0.5		0.030	0.998	1.45
	11	6.0	8.40	10	0.5		0.053	0.994	2.70
				N aO 3 S	N N
	12	6.0	8.40	10	0.2	H O	0.029	0.991	2.92
	13	6.0	8.40	10	0.8		0.057	0.969	2.10
	14	6.0	8.40	10	1.2		0.049	0.976	2.10
	Cathode						Anode

Table 7 -

 7 1 demonstrated that the increase of current density from 3.34 to 25.2 mA/cm2

  and appaent rate constants decreased to 0.049 min -1 (Table7-1). When Fe 3 O 4 particles were added in excess, an excess of sulfate radicals would be generated, resulting in considerable disappearance of sulfate radicals without the degradation of Orange II due to the combination between sulfate radicals themselves (Eq. (7-13)). Moreover, the interaction between excess Fe(II) on the surface of Fe 3 O 4 particles and sulfate radicals might directly quench sulfate radicals via Eq. (7-14)(Yan et al. 2011).As a result, the decomposition of PDS increased with an increase in Fe 3 O 4 dosage of up to 0.8 g/L (Table7-1), and the highest decolorization efficiency of Orange II was obtained at Fe 3 O 4 dosage of 0.8 g/L. This value was chosen as the optimal Fe 3 O 4 dosage and applied in the following experiments.

			1.0					3.4 mA/cm	2
			0.8					8.4 mA/cm 16.8 mA/cm	2 2
	SO 4	•-+ SO 4	0.6 •-→ 2SO 4	2-				25.2 mA/cm	2	(7-13)
	0.4 ≡Fe(II) + SO 4 A/A 0 •-→ ≡Fe(III) + SO 4	2-		(7-14)
			0.2
			0.0
			0	10	20	30	40	50	60
							time (min)

4

•-became more significant (Eq. (7-12))

Table 7 -

 7 2 GCMS identified reaction intermediates during EC/Fe 3 O 4 /PDS process

	Compound	Retention time (min)	Chemical name	Chemical structure
	A	6.03	1,4-benzoquinone	O	O
				O
				O
	B	16.02	1,2-naphthalenedione
				O
	C	12.75	1(3H)-isobenzofuranone	O
				O
	D	12.24	phthalic anhydride	O
				O
	E	12.26	methyl hydrogen phthalate

Table 7 -

 7 3 Comparison of electrical consumption under the different reaction system

Acknowledgements I would like to thank gratefully to my supervisors Prof. Mehmet A. Oturan and Prof. Hui Zhang for giving me the chance to conduct the thesis work. It is my pleasure to work with Prof. Mehmet A. Oturan and Prof. Hui Zhang and I learned a lot of things from them. I would also like to thank Dr. Nihal Oturan for giving me precious advice and technical supports.

I wish to express my gratitude to Prof. Ignacio Siré s (Barcelona University, Spain) and Prof. Zhihui Ai (Central China Normal University, China) for reading and evaluating my thesis.

I also wish to thank Dr Nihal Oturan (Université Paris-Est, France) and Professor Feng Wu (Wuhan University, China) for their acceptation to be part of thesis jury.

I also wish to thank all my colleagues in Advanced Oxidation Lab of Wuhan University

Chapter 2 Literature review

Filho, J.C., Santini, A.O., Nasser, A.L.M., Pezza, H.R., Eduardo de Oliveira, J., Melios, C.B. and Pezza, L. (2003). Potentiometric determination of saccharin in commercial artificial sweeteners using a silver electrode. Food Chemistry 83 [START_REF] Heng | Degradation of clofibric acid in aqueous solution by an EC/Fe 3+ /PMS process[END_REF], 297-301. Forgacs, E., Cserhá ti, T. and Oros, G. (2004) Removal of synthetic dyes from wastewaters: a review. Environment International 30(7), 953-971. Galletti, G.C., Chiavari, G. and Bocchini, P. (1995) Thermal decomposition products of aspartame as determined by pyrolysis-gas chromatography/mass spectrometry.

Journal of Analytical and Applied Pyrolysis 32, 137-151.

Gan, Z., Sun, H., Feng, B., Wang, R. and Zhang, Y. (2013) Occurrence of seven artificial sweeteners in the aquatic environment and precipitation of Tianjin, China.

Water Research 47( 14), 4928-4937. Ganesh, R., Boardman, G.D. and Michelsen, D. (1994) Fate of azo dyes in sludges.

Water Research 28 [START_REF] Heng Lin | Treatment of organic pollutants using electro-Fenton and electro-Fenton-like process in aqueous solution[END_REF], 1367-1376.

Garrido-Ramí rez, E.G., Mora, M.L., Marco, J.F. and Ureta-Zañartu, M.S. ( 2013 Pt was used as anode, a progressive accumulation of Cl  was observed, until reaching the expected maximum value of 0.6 mM. In contrast, the release of Cl-was increased significantly and reached a maximum concentration of 0.544 mM after 30 min, whereupon it decreased until a final value of 0.063 mM at 360 min. This phenomenon could be explained by the oxidation of Cl  to Cl 2 and/or ClO -/HClO, ClO 3 -, and ClO 4 by BDD( • OH) and/or • OH in the bulk [START_REF] Randazzo | Comparative electrochemical treatments of two chlorinated aliphatic hydrocarbons. Time course of the main reaction by-products[END_REF]Dirany et al. 2012).

The evolution of toxicity with reaction time

In order to monitored the potential toxicity of sucralose and its intermediates, 0.2 mM sucralose was electrolyzed in the existence of 0.2 mM Fe 2+ at 200 mA using Microtox method for both Pt/carbon-felt and BDD/carbon-felt cells. The percentage of bacteria luminescence inhibition versus the electrolysis time after 15 min exposure times of V. fischeri luminescent bacteria to the sucralose solutions was shown in Fig. 56.

When Pt was used as anode, there were several peaks due to the degradation primary intermediates and formation to secondary/tertiary intermediates that were more

The coefficients of determination R 2 gave the proportion of the total variation in the response variable explained. The R 2 obtained in this work (0.9507) ensured a satisfactory adjustment of the quadratic model to the experimental data. The coefficient of variation (CV) is a measure expressing standard deviation as a percentage of the mean. It illustrates the extent to which the data were dispersed. The lower values of CV give better reproducibility [START_REF] Ahmad | Optimization of coagulation-flocculation process for palm oil mill effluent using response surface methodology[END_REF][START_REF] Zinatizadeh | Process modeling and analysis of palm oil mill effluent treatment in an up-flow anaerobic sludge fixed film bioreactor using response surface methodology (RSM)[END_REF]. In this work, the CV value (10.6%) was within the acceptable range (0.5-13.5%) [START_REF] Zinatizadeh | Process modeling and analysis of palm oil mill effluent treatment in an up-flow anaerobic sludge fixed film bioreactor using response surface methodology (RSM)[END_REF]Wu et al. 2012a). In addition, adequate precision was used to assess the signal to noise ratio. The value of adequate precision greater than 4 is considered to give accurate prediction [START_REF] Körbahti | Response surface optimization of electrochemical treatment of textile dye wastewater[END_REF]. Therefore, the adequate precision value of 9.126 indicated adequate signals for the models to be used to navigate the design space. when PDS concentration was excessive [START_REF] Hori | Efficient decomposition of environmentally persistent perfluorocarboxylic acids by use of persulfate as a photochemical oxidant[END_REF][START_REF] Wu | Degradation of Acid Orange 7 in aqueous solution by a novel electro/Fe 2+ /peroxydisulfate process[END_REF].

Therefore, 10 mM was selected as the optimal initial PDS concentration and used in the following experiments. varied from 0.2 to 0.8 g/L, the decolorization efficiency of Orange II within 60 min reaction increased from 83.9% to 97.6% and apparent rate constants increased from 0.029 to 0.057 min -1 (Table 7-1). At higher Fe 3 O 4 dosage, more Fe(II) could be generated through Eq. (7-5), and then a considerable amount of sulfate radicals would be produced via Eq. (7-6). These sulfate radicals reacted immediately with Orange II, leading to a fast increase of the Orange II decolorization. When Fe 3 O 4 dosage further increased to 1.2 g/L, the Orange II decolorization efficiency declined slightly to 96.0% 7-2 and a plausible degradation pathway was proposed (Fig. 7-10) based on the results and previous studies (Özcan et al. 2009;Zhao et al. 2010a;Zhong et al. 2011b).

The cleavage of -N=Nof Orange II caused by electron transfer of sulfate radicals [START_REF] Yuan | Effects of chloride ion on degradation of Acid Orange 7 by sulfate radical-based advanced oxidation process: Implications for formation of chlorinated aromatic compounds[END_REF]) leads to the formation of sodium sulfanilamide and 1-amino-2-naphthol. Sodium sulfanilamide is thermally stable and highly soluble in water, whereas 1-amino-2-naphthol is oxygen sensitive and may be decomposed under aerobic conditions (Zhao et al. 2010a;Zhong et al. 2011b). Therefore, neither of these two intermediates was detected during GCMS analysis. Afterwards, sodium sulfanilamide was subsequently oxidized to 1,4-benzoquinone (A). 1-amino-2-naphthol was firstly oxidized to 1,2-naphthalenedione (B). 1,2-naphthalenedione (B) was subsequently oxidized to the quinoid structures, resulting in the formation of 1(3H)-isobenzofuranone (C), phthalic anhydride (D), methyl hydrogen phthalate (E) and 2-(hydroxymethyl) benzoic acid (F) (Özcan et al. 2009(Özcan et al. , Zhao et al. 2010a)) [START_REF] Heng | Degradation of C. I. Acid Orange 7 in aqueous solution by a novel electro/Fe 3 O 4 /PDS process[END_REF]50].

Further oxidation resulted in aromatic ring cleavage and generation of carboxylic acids and finally converted to CO 2 and H 2 O.

Chapter 8 General conclusion and future perspectives 8.1 General conclusion

In this paper, electro-Fenton and electro-Fenton-like processes were used to degrade artificial sweeteners and azo dyes. The results obtained during this thesis work concern the removal efficiency, the oxidation mechanism, degradation pathway and toxicity evolution of target pollutants.

(1) Electro-Fenton process was a effective method for the degradation of ASP in water. The oxidative degradation rate and mineralization efficiency were affected mainly by the catalyst (Fe 2+ ) concentration and applied current. The absolute rate constant of hydroxylation reaction of ASP was found as to be (5.23 ± 0.02) × 10 9 M -1 s - 1 . Short-chain aliphatic acids such as oxalic, oxamic and maleic acid were identified as aliphatic intermediates in the electro-Fenton process. The bacteria luminescence inhibition showed the toxicity of ASP solution decreased, during the treatment, after reached a maximum during the first period of the oxidation reaction.

(2) Artificial sweetener SAC could be degraded effectively by electro-Fenton process with a DSA, Pt or BDD anode. However, the using of BDD anode could accelerate the mineralization of SAC. The optimal conditions for SAC removal were SAC concentration 0.2 mM, Fe 2+ concentration 0.2 mM, Na 2 SO 4 concentration 50 mM, applied current 200 mA and initial pH 3.0. Oxalic, formic, and maleic acids were observed as aliphatic by-products of SAC during electro-Fenton process. The bacteria luminescence inhibition showed the toxicity of SAC solution increased at the beginning of electrolysis, and then it declined until the end of the reaction.

(3) Artificial sweetener SUC could be completely mineralized in a 360 min reaction by electro-Fenton process with a Pt or BDD anode. The mineralization rate was affected mainly by the Fe 2+ concentration and applied current. The mineralization current efficiency (MCE) decreased with rising applied current from 100 to 500 mA with both Pt and BDD anodes. Oxalic, pyruvic, formic and glycolic acids were detected during the oxidation of SUC.

(4) Orange II was effectively decolorized by EC/α-FeOOH/PDS process. The initial pH of Orange II solution had little effect on the decolorization of Orange II. RSM attacking organic compounds, hydroxyl radicals are more likely to do through hydrogen abstraction or addition reactions, while sulfate radicals participate in electron transfer reaction. Therefore, further investigation can focus on the different by-products generated under the attack of these two radicals.

(4) All the solutions contained organic pollutants are synthetic wastewaters which were prepared in the lab. In the future, we can use eletro-Fenton and sulfate radical-based electro-Fenton-like processes to degrade target pollutants in actural wastewater.