Par Wang

PR Weihua Song

MCF Corinne Ferronato

Khalil Hanna

devant la commission d'examen

Keywords: Chimie théorique, physique, analytique Chimie théorique, physique, analytique pharmaceutical, advanced oxidation processes, photocatalytic degradation, heterogeneous, homogeneous, Photo-Fenton, BiOCl0.75I0.25, Acetaminophen, p-hydroxyphenylacetic acid

In recent decades, pharmaceuticals and personal care products (PPCPs) have emerged as a new environmental concern. Pharmaceuticals have been widely detected in ground water, surface water and even drinking water with a concentration ranging from ng to µg L -1 in wastewater effluents. Sewage treatment plants (STPs) cannot remove completely all pharmaceuticals due to their low concentration in water and resistance to biological degradation. Levels of many pharmaceutically active compounds (PhACs) are barely reduced and can be detected in wastewater-treatment plant (WWTP) effluents. In order to overcome this problem, advanced oxidation processes (AOPs) usually are adopted for the application of this wastewater treatment, which can be attributed to the strong oxidation abilities of different radicals produced.

In AOPs, the most frequently used methods include homogeneous reaction such as Fenton, Fenton like or Photo-Fenton process. Another one is heterogeneous reactions involving semiconductors to perform also called catalytic degradations.

During these reactions, ultraviolet (UV) and visible lights are generally used as activation source. In this work, simulated solar light (300 nm<𝝀<500 nm) was taken

as radiation source to perform heterogeneous and homogeneous AOPs to efficiently degradate pharmaceutical. This results was obtained producing different kind of radicals including hydroxyl (HO • ), sulfate (SO4 •-) and hydroperoxide/superoxide (HO2 • /O2 •-) radicals.

Acetaminophen (ACTP) and p-hydroxyphenylacetic acid (p-HPA) were two typical pharmaceuticals. As one kind of the most commonly detected pollutant, ACTP is extensively used globally as an analgesic and antipyretic drug, and p-HPA is one of the pharmaceutical intermediates, widely used in the synthetisation of pesticides and commonly detected in olive oil wastewater. In this work, ACTP and p-HPA are taken as the representative compounds of pharmaceutical pollutant.

In the first part of my research, composite catalyst BiOCl0.75I0.25 was synthesized by precipitation method under ambient pressure and low temperature followed with a series of characterization. Catalyst showed flower-like microspheres and possessed great surface area. Most importantly, from the detection of UV-vis diffuse reflection spectra (DRS), the band gap of BiOCl0.75I0.25 was 2.12 eV, was determined. Moreover, the adsorption and photocatalytic degradation abilities of bismuth catalyst BiOCl0.75I0.25 in water are tested under simulated solar light using p-HPA as chemical pollutant. On the basis of strong adsorption of p-HPA on the surface of catalyst, this section of my work aims at investigating the interfacial mechanism in the heterogeneous system through adsorption/desorption experiments using competition desorption with phosphate anions. Key factors affecting the degradation kinetic of p-HPA such as pH and dissolved oxygen are investigated. It is found that solution pH has significant effect on the adsorption and degradation rate, depending on the zeta potential of catalyst and molecular form of p-HPA. p-HPA shows a maximum adsorption at pH 4.5 and faster degradation rate at pH 3.0. Dissolved oxygen concentration is a key parameter for the pollutant removal through formation of hydroperoxide/superoxide radical anion couple (HO2 • /O2 •-). These radicals, that are the main reactive species involved in the reaction occurring after irradiation of catalyst, are detected using competition kinetic approach and selective radical probes.

In the second part, p-HPA was taken as representative pollutant. Our group has done a lot of work on a new complexing agent of iron, the Ethylenediamine-N,N'-disuccinic acid (EDDS). EDDS is a structural isomer of EDTA, and it's readily biodegradable. It has been proposed as a safe and environmentally benign replacement for EDTA for environmental remediation products. Our laboratory has studied the physicochemical properties of the Fe(III)-EDDS complex, and its application in the Fenton, Photo-Fenton and persulfate (PS) activation as iron source. While the comparison between Fe(III)-EDDS/H2O2/UV and Fe(III)-EDDS/PS/UV systems under same conditions was studied for the first time including the effects of pH, Fe(III)-EDDS concentration, H2O2 and PS concentrations in this part. Results showed that Fe(III)-EDDS/H2O2/UV showed higher degradation efficiency than Fe(III)-EDDS/PS/UV, especially in neutral and alkaline solutions. In Fe(III)-EDDS/H2O2/UV reaction, p-HPA degradation efficiency increased fast from pH 2.5 to 7.5, then it began to decrease when pH increased to 9.0. While degradation rate (Rp-HPA) started to decrease with pH increase to 3.9 in Fe(III)-EDDS/Na2S2O8/UV.

The degradation rate of p-HPA increased using higher concentrated solution of H2O2 and Na2S2O8. However, increasing the Fe(III)-EDDS concentration, the degradation efficiency of degradation decreases. The second order constant of p-HPA with HO  or SO4 •-radicals under different pH were determined by laser flash photolysis (LFP) experiments for the first time. Results showed that kp-HPA,HO• was higher than kp-HPA,SO4•-for both anionic and molecular forms of pollutant.

Finally, different Iron-based homogeneous and heterogeneous AOPs were tested on the ACTP degradation. Moreover, reactivity of selected pollutant with HO  , SO4 •- and HO2 • /O2 •-that are produced by UV/H2O2, UV/Na2S2O8, Fe(III)-EDDS/H2O2/UV and Fe(III)-EDDS/PS/UV system was investigated. The effects on the degradation of pH and chloride ions were studied at the same time. Results indicate that using UV/Na2S2O8 system, a better removal efficiency (57%) was obtained compared to UV/H2O2 (11%). UV/H2O2 showed best degradation efficiency in neutral solution, while using UV/Na2S2O8 the highest removal rate was observed in alkaline solutions.

Chlorine ions (10 mM) has negligible effects in these two systems. However, the same concentrated iodine ions can accelerate the degradation in the dark (Fenton-like reaction). Fe(III)-EDDS/H2O2/UV and Fe(III)-EDDS/PS/UV showed a much faster degradation efficiency than UV/H2O2 and UV/Na2S2O8 using the same irradiation source. On the contrary, Fe(III)-EDDS/H2O2/UV was more efficient than Fe(III)-EDDS/PS/UV, especially in neutral and basic solution.

Hydroperoxide/superoxide radicals (HO2 • /O2 •-) are produced by the UV/BiOCl0.75I0.25 photocatalytic process, which exhibited higher degradation rate than UV/H2O2 and UV/Na2S2O8. ACTP can be completely removed in two hours at pH 3.

Interfacial mechanism investigation showed that reaction mainly occurred on the surface of catalyst or in the bulk solution that closed to the surface of catalyst. Second order constant between ACTP and HO  or SO4 •-radicals were determined by LFP and competition method, respectively.

Introduction

Pharmaceuticals in water have emerged as one of the most urgent issues due to their massive use and potential chronic health effects. Pharmaceutically active compounds (PhACs) have been detected in sewage treatment plant, surface water and even drinking water at a concentration of ng to µg per litre. Traditional wastewater treatment plant cannot remove all the pollutant compound due to its trace concentration and resistance to micro-biology degradation. So advanced oxidation processes (AOPs) can be successfully used in the field of wastewater, underground water and gas treatment, to convert toxic and bio-recalcitrant contaminants into biodegradable by-products, to remove color or to reach the complete mineralization of organic pollutants. Based on the high oxidation abilities of radicals generated in AOPs, AOPs showed a promising potential in practical use. Usually, radicals originated from the activation of H2O2, Na2S2O8 or O3 etc. by ultraviolet (UV)-visible light but also using ferrous, ferric ions, ultrasonic or semiconductor catalyst.

Nowadays, the most commonly used AOPs are homogeneous reactions using iron as activation species (Fenton, Photo-Fenton, Fenton-like). Between the heterogeneous processes photo-degradation processes involving semiconductor catalysts are widely adopted. Numerous research papers focus on the synthesization, characterization and pollutants degradation using different catalyst. Among these catalyst, TiO2 has attracted the most attention due to its stable physico-chemical properties, non-toxicity and high degradation efficiency. However, the large band gap (3.2 eV) of TiO2 restricts its use under a small range of solar light; therefore, a semiconductor photocatalyst which could make full use of solar light would be a major progression. Bismuth-based compounds such as Bi2O3, Bi2O2CO3, Bi2S3 and BiOX (X = Cl, Br, I) have received wide attention due to their characteristic hierarchical structures and high photocatalytic activities. In particular, BiOX with a layered structure has been reported to have internal electric fields between [Bi2O2] and [Cl2], which are favourable for the efficient photo-induced electron-hole separation. Nowadays, majority of BiOX researches focused on the synthesization, modification, composite or doping method to increase its photo-catalytic activities.

While mechanism of the interfacial system such as chemical reactions, adsorption kinetics and influence of solution factors are not deeply investigated.

In homogeneous AOPs, Fenton, Photo-Fenton and activation of sulfate radicals are research hot topics. Most of these reactions are based on the activation of Fe(II)/Fe(III) using UV irradiation. However, under neutral pH conditions a low efficiency of pollutant degradation system is observed. This low efficiency is mainly due to the presence of insoluble forms of Fe(III) under neutral pH conditions. So, in our group, we investigated the properties of new complexing agent Ethylenediamine-N,N'-disuccinic acid (EDDS), which is a structural isomer of Ethylenediaminetetraacetic acid (EDTA). It has been proposed as a safe and environmentally benign replacement for EDTA for environmental remediation products as it is also a strong complexing agent. Fe(III) is complexed by EDDS with a ratio 1:1, the physicochemical properties of Fe(III)-EDDS complex have been studied in detail. Fe(III)-EDDS complex has higher quantum yield of  OH radical formation at higher pH values. Furthermore, Fe(III)-EDDS complex showed an excellent activation efficiency in Photo-Fenton, Fenton and Na2S2O8 process as iron source in our previous research. However, comparison between activation of H2O2 and Na2S2O8 under same conditions has not been studied before.

Acetaminophen (ACTP) is a pharmaceutical with a heavy use and frequent detection in global. p-hydroxyphenylacetic acid (p-HPA) a pharmaceutical intermediate for the synthesis such as Atenolol, and it was also commonly detected in olive oil wastewater. ACTP and p-HPA are taken as the representative compounds of pharmaceutical pollution.

The aim of my thesis is to investigate the degradation of ACTP and p-HPA by different AOPs under simulated solar light, including homogeneous and heterogeneous reactions. In heterogeneous reactions, BiOCl0.75I0.25 was taken as the catalyst ascribing to its good response to solar light. The most important attention will focus on the interfacial mechanism of the catalyst induced reactions. In homogeneous reactions, Fe(III)-EDDS complex is taken as iron source to activate H2O2 and Na2S2O8 under same conditions. Laser flash photolysis was adopted to check the second order constant between compound with HO  and SO4 •-. Simulated solar light (300 nm<𝝀<500 nm) was taken as irradiation during the whole research.

In the first part, p-HPA was taken as model compound, because p-HPA showed a much stronger adsorption on the surface of catalyst BiOCl0.75I0.25 than ACTP.

Composite catalyst BiOCl0.75I0.25 was synthesized in EG-H2O mixed solvent by precipitation method, and our previous studies has proved that BiOCl0.75I0.25 was a promising candidate in the solar-light driven photo-degradation. In this part, followed with the synthesis and characterization of BiOCl0.75I0.25, research about active species identification, pH effects, interfacial mechanism, oxygen effects, synergistic reaction with Na2S2O8 and mineralization in the photo-catalytic degradation of p-HPA were investigated in detail.

In the second part, p-HPA was also taken as the representative compound.

Compare of p-HPA degradation rate (Rp-HPA) in Fe(III)-EDDS/H2O2/UV and Fe(III)-EDDS/Na2S2O8/UV systems was evaluated under same conditions for the first time. The effects of UV, H2O2 and Na2S2O8 concentration, Fe(III)-EDDS concentration and pH were studied seriously. Furthermore, using laser flash photolysis (LFP), second order constants between p-HPA and HO  , SO4 •-were determined under different pH.

Finally, ACTP was taken as the representative compound. We compared the degradation efficiency by different AOPs. ACTP degradation in UV/H2O2, UV/Na2S2O8, Fe(III)-EDDS/H2O2/UV, Fe(III)-EDDS/Na2S2O8/UV and UV/ BiOCl0.75I0.25 were investigated. Effects of pH, halogen ions and interfacial mechanism were studied at the same time. Second order constant between ACTP and HO  , SO4 •-was determined by LFP and competition method, respectively.

Bibliography

2-1. Pharmaceutical pollution

With the fast economic development, a better living condition leads to longer life expectancy, which increased the total population, in particular the elderly group. It may result in increase in the demand of pharmaceutical for people in domestic use or in hospital 1 . In recent decades, pharmaceuticals and personal care products (PPCPs)

have emerged as a new environmental concern due to its wide spread, huge amount and undetectable biotoxicity. Pharmaceuticals have been widely detected in ground water, surface water and even drinking water. These kinds of pollutants include analgesics, anti-inflammatory drugs, antibiotics, lipid regulators, psychiatric drugs and β-blockers, detected in concentrations ranging from nanograms per litre to micrograms per litre in wastewater influent. Sewage treatment plants (STPs) cannot remove all kinds of pharmaceuticals completely due to their trace concentration in water and resistance to biological degradation. Levels of many pharmaceutically active compounds (PhACs) are barely reduced and can be detected in wastewater-treatment plant (WWTP) effluents. Although the amount of these pharmaceuticals in the aquatic environment is low, its continuous input may constitute in the long-term a potential risk for aquatic and terrestrial organism. Therefore, over the past few years they are considered to be an emerging environmental problem [2][3][4][5][6] .

2-1-1. Source

Normally, the source of pharmaceuticals presented in water can be divided into two ways: point source pollution and diffuse source pollution. Point source pollutant is a single identifiable source which originates from separate locations and can be calculated in mathematical modeling like pharmaceutical industrials or hospitals. On the contrary, diffuse source pollutant is hard to be identified the discrete location which occurs over board geographical scales like rivers in countryside. Fig. 2-1 reports the sources and receiving media of pharmaceutical pollution. There are six major sources namely, landfill, animal waste, freshwater aquaculture waste, hospital waste, industrial waste and domestic waste which is summarized and simplified by several review literatures. The receptors of pharmaceuticals can be divided as three major bodied in the natural environment, for instance, soil zone, groundwater and surface water 1,7 .

Fig.2-1

Potential sources and pathways of pharmaceutical pollution in soil and water

2-1-2. Distribution and detection

In addition to the pharmaceutical industry, the effluent from the wastewater treatment plant also contains a large number of pharmaceutical compounds.

Aleksandra Jelic et al. 8 collected 72 samples in two years to study the residues of medical materials in the influent, effluent and sediment of three traditional sewage treatment plants. These samples belong to different kinds of medicine, including non-steroidal anti-inflammatory drugs, lipid regulators (fibrates and statins), psychotropic drugs (benzodiazepines and antiepileptic drugs), histamine H2-receptor antagonist, systemic use of antibacterial agents, β-blockers, β-agonists, diuretics, angiotensin angiotensin converting enzyme (ACE) inhibitors and anti-diabetic drugs.

The results showed that in the influent and effluent 32 kinds and 29 different organic molecules were respectively detected and their concentration was in the range of µg -ng L -1 . 21 kinds of pharmaceuticals were detected in the sediment of the three wastewater treatment plant, with a concentration of about 100 ng g -1 . These pollutants show different removal efficiencies, even if they belong to the same type of drug.

Jelena et al. 9,[START_REF] Radjenovic | Fate and distribution of pharmaceuticals in wastewater and sewage sludge of the conventional activated sludge (CAS) and advanced membrane bioreactor (MBR) treatment[END_REF] studied the concentration of 31 active substances in sewage and solid phase after treated by traditional activated sludge process (CASs) and membrane bioreactor (MBRs). These drugs belong to different species and have different physical and chemical properties. In these two kinds of treatment methods, the most commonly detected pollutants are analgesic and anti-inflammatory drugs, including ibuprofen (14.6-31.3 µg L -1 ) and acetaminophen (7.1-11.4 µg L -1 ), antibiotic ofloxacin (0.89-31.7 µg L -1 ), lipid regulators gemfibrozil (2.0-5.9 µg L -1 ) and bezafibrate (1.9-29.8 µg L -1 ) β-blocker atenolol (0.84-2.8 µg L -1 ), hypoglycemic agent glibenclamide (0.12-15.9 µg L -1 ) and diuretic hydrochlorothiazide (2.3-4.8 µg L -1 ). In addition, several drugs such as ibuprofen, ketoprofen, diclofenac, ofloxacin and azithromycin were detected in the sludge, at concentration of 741. 1, 336.3, 380.7, 454.7 and 299.6 ng g -1 dry weight, respectively.

According to literature data, the most commonly detected drugs in soil include species reported in Table 2-1. Generally, the concentration of drugs in soil is lower than that in water. Studies have shown that in China five kinds of the six most common drugs are regularly detected. However, the detection of antibiotics is the world's highest concentration, followed by Kuwait and the United States. 4 The detection of drugs and active substances in water is shown in Table 2-2, which listed several frequently detected pollutants in plant effluent, surface water and groundwater. The drug concentration was higher in the sewage treatment plant, followed by the surface water and fresh water, mainly because of natural causes in the water, for example by biotransformation, photolysis, dispersion, adsorption and volatilization etc [START_REF] Gurr | Harnessing natural attenuation of pharmaceuticals and hormones in rivers[END_REF] . In Asia, the most commonly detected pollutants in effluent and fresh water are antibiotics, such as trimethoprim and sulfamethoxazole a: undetectable

2-1-3. Degradation

Recently, the degradation of pharmaceutical pollutant included various physical, chemical and microbiology methods 5,[START_REF] Zhang | Study of a hybrid process combining PACs and membrane operations for antibiotic wastewater treatment[END_REF][START_REF] Chen | A novel application of TPAD-MBR system to the pilot treatment of chemical synthesis-based pharmaceutical wastewater[END_REF][START_REF] Ng | A novel application of anaerobic bio-entrapped membrane reactor for the treatment of chemical synthesis-based pharmaceutical wastewater[END_REF][START_REF] Mendez-Arriaga | Photocatalytic degradation of non-steroidal anti-inflammatory drugs with TiO2 and simulated solar irradiation[END_REF][START_REF] Taheran | Degradation of chlortetracycline using immobilized laccase on Polyacrylonitrile-biochar composite nanofibrous membrane[END_REF][START_REF] Wu | Visible-light-assisted peroxymonosulfate activation and mechanism for the degradation of pharmaceuticals over pyridyl-functionalized graphitic carbon nitride coordinated with iron phthalocyanine[END_REF][START_REF] Zhang | Degradation kinetics and mechanism of sulfadiazine and sulfamethoxazole in an agricultural soil system with manure application[END_REF][START_REF] Adityosulindro | Sonolysis and sono-Fenton oxidation for removal of ibuprofen in (waste) water[END_REF][START_REF] Ben | Oxidation of sulfonamide antibiotics by chlorine dioxide in water: Kinetics and reaction pathways[END_REF][START_REF] Deng | One-step in situ hydrothermal fabrication of octahedral CdS/SnIn4S8 nano-heterojunction for highly efficient photocatalytic treatment of nitrophenol and real pharmaceutical wastewater[END_REF][START_REF] Nomura | Removal behaviors of sulfamonomethoxine and its degradation intermediates in fresh aquaculture wastewater using zeolite/TiO2 composites[END_REF][START_REF] Savun-Hekimoglu | Decomposition of PPCPs by ultrasound-assisted advanced Fenton reaction: A case study with salicylic acid[END_REF][START_REF] Daoud | Ozone pretreatment of process waste water generated in course of fluoroquinolone production[END_REF][START_REF] Hussain | Photo-assisted electrochemical degradation of sulfamethoxazole using a Ti/Ru0.3Ti0.7O2 anode: Mechanistic and kinetic features of the process[END_REF][START_REF] Li | Heterogeneous catalytic ozonation of clofibric acid using Ce/MCM-48: Preparation, reaction mechanism, comparison with Ce/MCM-41[END_REF][START_REF] Serna-Galvis | Degradation of highly consumed fluoroquinolones, penicillins and cephalosporins in distilled water and simulated hospital wastewater by UV254 and UV254/persulfate processes[END_REF][START_REF] Hofman-Caris | Removal of pharmaceuticals from WWTP effluent: Removal of EfOM followed by advanced oxidation[END_REF] , and it can be divided into these classifications:

1) Microbiological degradation, which is used in many domestic and foreign wastewater treatment plants, including activated sludge process and membrane bioreactor method [START_REF] Radjenovic | Analysis of pharmaceuticals in wastewater and removal using a membrane bioreactor[END_REF] . The treatment of pharmaceutical wastewater by microbiological treatment is characterized by low cost and avoiding the second pollution. At the same time, there are also some disadvantages such as leaving residue in effluent and sludge.

Jelena et al. [START_REF] Radjenovic | Analysis of pharmaceuticals in wastewater and removal using a membrane bioreactor[END_REF] studied the degradation of different classes of pharmaceuticals, such as analgesics, anti-inflammatory agents, lipid modulators and antibiotics in a laboratory membrane bioreactor, and compared with the traditional activated sludge process. The membrane bioreactor showed a stronger removal efficiency of medical pollutants than the activated sludge process. For most of the drugs, the removal rate can achieve 80%, and the drug concentration in the effluent is also more stable than the activated sludge process. For Carbamazepine, which is the most resistant biodegradable compound, it cannot be removed or transformed by membrane bioreactor and activated sludge method, because the COD, TOC and ammonia nitrogen of effluent have not been greatly improved.

2) Adsorption, representing the treatment of wastewater by adsorption method predominantly depended on the physical or chemical adsorption actions between the compounds and adsorbent such as activated carbon, and the characterizations showed below [START_REF] Yamamoto | In-situ Adsorption of Polymer Particles on Multi-wall Carbon Nanotubes Using Colloidal Techniques[END_REF][START_REF] Bahamon | Computational study of ibuprofen removal from water by adsorption in realistic activated carbons[END_REF][START_REF] Álvarez-Torrellas | Effective adsorption of non-biodegradable pharmaceuticals from hospital wastewater with different carbon materials[END_REF] :

( iron, copper, nickel, chromium, zinc, cobalt; (5) The saturated carbon can be reused after regeneration, avoiding the second pollution. A lot of research and practice have proved that activated carbon is a kind of excellent adsorbent. It has special treatment effect in the treatment of industrial wastewater.

2-2. Model compounds

2-2-1. p-Hydroxyphenylacetic acid (p-HPA): presence and degradation

p-hydroxyphenylacetic acid is one a phenolic pollutant, mainly from the agricultural industrial plant, such as olive oil wastewater and distilleries, and so on [START_REF] Benitez | The role of hydroxyl radicals for the decomposition of p-hydroxy phenylacetic acid in aqueous solutions[END_REF] .

Its chemical structure is shown in Fig. 2-2, and physical chemical properties are reported in Table 2-3. In addition, as one kind of organic intermediates for the synthesis of the production of β-blocker atenolol, 4,7-two hydroxy isoflavone and pesticide, it is currently regarded as one popular intermediate favored by the majority of medical experts. While with a large amount of p-hydroxyphenylacetic acid discharging into the water, this phenolic compound will cause a series of damage to the water environment. Although the environmental harm of p-hydroxyphenylacetic acid is not reported, a lot of discharge will result in unknown dangers in the future.

Benitez et al [START_REF] Benitez | The role of hydroxyl radicals for the decomposition of p-hydroxy phenylacetic acid in aqueous solutions[END_REF] ivestigated the reaction between p-HPA and hydroxyl radical (HO  ), which was produced by UV/H2O2 and Fenton process. Comparing with the direct photolysis under UV irradiation, the presence of H2O2 accelerates the degradation efficiency. At the same time, the effects of Fe 2+ concentration H2O2 concentration and pH of solution were discussed in detail. Finally, the second order constant between p-HPA and HO  was determined to be 7.02×10 8 M -1 s -1 . Miguel A.

Miranda [START_REF] Miranda | Pyrylium salt-photosensitized degradation of phenolic contaminants present in olive oil wastewater with solar light: Part III. Tyrosol and p-hydroxyphenylacetic acid[END_REF] Furthermore, the degradation pathway and intermediates of these two compound were elucidated showing the formation of different aromatic compound and short chain organic acid through hydroxylation and decarboxylation. Electro chemical oxidation methods were adopted by Flores et al. [START_REF] Flores | Removal of 4-hydroxyphenylacetic acid from aqueous medium by electrochemical oxidation with a BDD anode: Mineralization, kinetics and oxidation products[END_REF] to investigate the degradation of p-HPA. In this work, mineralization, kinetic reaction and photoproducts were reported. The TOC removal and degradation of p-HPA both followed the pseudo first order kinetic reactions. Finally, p-hydroxybenzaldehyde,p-hydroxybenzyl alcohol and acetic acid were the main intermediates during the process. However no significant decrease of TOC was reported. Flores et al. [START_REF] Flores | 4-Hydroxyphenylacetic acid oxidation in sulfate and real olive oil mill wastewater by electrochemical advanced processes with a boron-doped diamond anode[END_REF] investigated the degradation of p-HPA by HO  produced by H2O2, electro-Fenton and photo-electro-Fenton, and excellent mineralization was achieved.

2-2-2. Acetaminophen (ACTP): presence and degradation

Acetaminophen (ACTP), that is a drug and personal care product (PPCPs), is used as an alternative to aspirin, which mainly plays a role in relieving pain. Its chemical structure is shown in Fig. 23, and physical chemical properties are reported in Table 23. The global consumption of ACTP is estimated to be about 70 thousand tons, and estimated the existing production capacity of 90 thousand tons. As for the huge quantity of production and consumption, acetaminophen has become an important object in environmental treatment studies. Younghee Kim [START_REF] Kim | Aquatic toxicity of acetaminophen, carbamazepine, cimetidine, diltiazem and six major sulfonamides, and their potential ecological risks in Korea[END_REF] investigated the toxicity of ACTP on one kind of Marine bacteria, freshwater invertebrates, and Japanese green fish. Results showed that ACTP presents a risk for these organisms when its concentration increased beyond 1.8 ug L -1 . In Spain although the removal efficiency of ACTP in wastewater treatment plants can achieve 90%, it also can be detected in surface water with a concentration 0.033 µg L -1 , and 0.22 µg L -1 in effluent of sewage treatment plant (STP) [START_REF] Tan | Kinetic study of acetaminophen degradation by UV-based advanced oxidation processes[END_REF] . In South Korea, ACTP was detected with a concentration 0.071 and 6.8 µg L -1 in surface water and effluent of STP, respectively.

Moreover, in the process of chlorine disinfection, ACTP can be transformed to more toxic substances. So acetaminophen has been regarded as a potential environmental pollutant by environmental scientists.

Recently, numerous researches focus on the degradation of ACTP by advanced oxidation processes (AOPs). Zhang et al. [START_REF] Zhang | Photodegradation of acetaminophen in TiO(2) suspended solution[END_REF] investigated the photocatalytic degradation of ACTP by TiO2 under irradiation. TiO2 based photocatalytic degradation can efficiently remove ACTP from water. Under pH 3, with the addition of TiO2 0.5 g L -1 , the degradation of ACTP can obtain 59.1% with an initial concentration 30 µmol L -1 after 60 min. Andreozzi et al. [START_REF] Andreozzi | Paracetamol oxidation from aqueous solutions by means of ozonation and H2O2/UV system[END_REF] degraded ACTP by using ozone and H2O2 in aqueous solution. These two methods both can attack the benzene rings and the mineralization can achieve 30% and 40% respectively under the current conditions. Furthermore, the intermediates of ACTP were identified by GC-MS. Mark Daniel G. de Luna et al. [START_REF] De Luna | Acetaminophen degradation by electro-Fenton and photoelectro-Fenton using a double cathode electrochemical cell[END_REF][START_REF] De Luna | Kinetics of acetaminophen degradation by Fenton oxidation in a fluidized-bed reactor[END_REF] investigated the decomposition of ACTP by Fenton process regarding the dosage of ferrous and influences of [Fe 2+ ]/[H2O2] ratio. Main intermediates were identified by SPME/GC-MS. Chia-Chi Su et al. [START_REF] Su | Degradation of acetaminophen in an aerated Fenton reactor[END_REF] discussed on the degradation of ACTP by aeration Fenton. Solution pH, concentration of Fe 2+ and H2O2 were studied on the effects on the removal efficiency and mineralization was reported. Alam et al. [START_REF] Trovo | Paracetamol degradation intermediates and toxicity during photo-Fenton treatment using different iron species[END_REF] studied the degradation of ACTP in photo-Fenton processes by taking FeSO4 and iron complex as different iron source under solar light irradiation. Results demonstrate a significant degradation in the presence of FeSO4 and same intermediates (with two different sources of iron) were identified. Meanwhile, the toxicity assessment using vibrio fischeri and water flea showed that these two different photo-Fenton processes can decrease the toxicity of solution to 40% representing a safe and effective method for the degradation of ACTP. oxidation processes (AOPs) present particular advantages for the degradation of persistent organic pollutant (POPs) up to the trace level concentration. [START_REF] Yang | Efficient removal of organic contaminants by a visible light driven photocatalyst Sr6Bi2O9[END_REF][START_REF] Pignatello | Advanced oxidation processes for organic contaminant destruction based on the Fenton reaction and related chemistry[END_REF] As one new and effective pollutant control technics, AOPs emerged in history from 1980s.

AOPs efficiency are based on the formation of various radicals by semiconductor catalyst, irradiation, electricity, ultrasonic and activation of H2O2, Na2S2O8, and ozone.

H2O2 alone is a weak oxidation with E 0 (H2O2/H2O) = 1.763 V/SHE in acidic solution and E 0 (H2O2/OH -) = 0.88 V/SHE in alkaline medium. It's only effective for attacking reduced sulfur compound, cyanides and certain organics such as aldehydes, formic acid and some nitro-organic and sulfo-organic compound. Its oxidation power can be notably improved by combination with ozone, UV radiation and transition metals such as iron ions. The latter technology is the so-called Fenton chemistry explained in the section below. Among the reactive species involved in AOPs, the most commonly researched radicals include HO • (E 0 =2.80V) [START_REF] Sherwood | Modified Fenton oxidation of diesel fuel in arctic soils rich in organic matter and iron[END_REF][START_REF] Giannakis | Effect of Fe(II)/Fe(III) species, pH, irradiance and bacterial presence on viral inactivation in wastewater by the photo-Fenton process: Kinetic modeling and mechanistic interpretation[END_REF] , SO4 •-(2.6-3.2 V) [START_REF] Wu | Sulfate Radical Photogeneration Using Fe-EDDS: Influence of Critical Parameters and Naturally Occurring Scavengers[END_REF][START_REF] Parsons | Advanced oxidation processes for water and wastewater treatment[END_REF][START_REF] Nie | Degradation of chloramphenicol by persulfate activated by Fe2+ and zerovalent iron[END_REF][START_REF] Nie | Degradation of chloramphenicol by thermally activated persulfate in aqueous solution[END_REF] . All of the radicals formed by AOPs possess high oxidation abilities, and standard reduction potential in aqueous medium of the most commonly reported oxidizing agents was showed in the following Table2-4.

For the source of  OH radical in AOPs, a combination of conventional strong oxidants such as ozone, oxygen or hydrogen peroxide and catalysts (e.g., transition metal, iron and copper), semiconductor solid (e.g., TiO2 and Fe2O3) together with energy sources of UV-visible irradiation, ultrasound, electricity or ionizing radiation was used. Once the radicals are formed, it will induce a series of free radical chain reactions, which can make most of the organic compounds mineralized or decomposed into low toxic or non-toxic small molecules, and even directly degrade to carbon dioxide and water. The pharmaceutical pollutant in water cannot be completely removed by microbiological method. It can be treated by advanced oxidation process for the second time, so as to reduce its toxicity and achieve discharge standard [START_REF] Huang | Combination of Fenton processes and biotreatment for wastewater treatment and soil remediation[END_REF] .

Among the activation methods in AOPs, light is a kind of green and sustainable energy, the application of light irradiation into wastewater treatment is of great interest in related research field. The basic and popular systems of this kind, excluding the application of iron, include UV/H2O2 [START_REF] Li | The roles of halides in the acetaminophen degradation by UV/H2O2 treatment: Kinetics, mechanisms, and products analysis[END_REF] , UV/Na2S2O8 74 , UV/O3 [START_REF] Andreozzi | Ozonation and H2O2/UV treatment of clofibric acid in water: a kinetic investigation[END_REF][START_REF] Hua | Ozone treatment and the depletion of detectable pharmaceuticals and atrazine herbicide in drinking water sourced from the upper Detroit River, Ontario, Canada[END_REF] , UV/semiconductor [START_REF] Xiong | Degradation of acetaminophen by UVA/LED/TiO2 process[END_REF][START_REF] Doll | Photocatalytic degradation of carbamazepine, clofibric acid and iomeprol with P25 and Hombikat UV100 in the presence of natural organic matter (NOM) and other organic water constituents[END_REF][START_REF] Perez-Estrada | Decomposition of diclofenac by solar driven photocatalysis at pilot plant scale[END_REF] and photoelectrochemical [START_REF] Valdez | Degradation of paracetamol by advance oxidation processes using modified reticulated vitreous carbon electrodes with TiO2 and CuO/TiO2/Al2O3[END_REF][START_REF] Ghasemian | Photoelectrocatalytic degradation of pharmaceutical carbamazepine using Sb-doped Sn80%-W20%-oxide electrodes[END_REF] processes. For one example, the UV/H2O2 oxidizing ability can be attributed to the formation of  OH, O2 -and HO2  , the reaction process is shown as following [START_REF] Zhang | Kinetics, degradation pathway and reaction mechanism of advanced oxidation of 4-nitrophenol in water by a UV/H2O2 process[END_REF] :

Initiation:

H2O2 + hν → 2  OH (R2-1) H2O2 ↔ HO2 -+H + (R2-2)
Propagation:

 OH + H2O2 → HO2  + H2O (R2-3) H2O2 + HO2  → O2 + H2O +  OH (R2-4)

Termination:

 OH + HO2  →H2O + O2 (R2-5) With strong oxidation abilities, the free radicals can transform most organic pollutants degradation into biodegradable compound, or even directly oxidize them to carbon dioxide and water in aqueous solution. Because the position of the valence band and the conduction band is different, the photocatalytic activity of different semiconductors is different, and the degradation effect of the specific organic matter is also obviously different. Among them, TiO2 has the characteristics of high chemical stability, light resistant corrosion, harmless to human body, low cost, and deep valence band level. It is widely used in the field of pollutants degradation, air purification, decomposition of water, solar energy conversion and storage and other fields [START_REF] Shayegan | TiO2 photocatalyst for removal of volatile organic compounds in gas phase -A review[END_REF] . However, TiO2 photocatalytic technology has some shortcomings, which restrict its wide application. First of all, TiO2 possesses large band gap (rutile type TiO2 band gap is 3.0 eV, anatase type TiO2 band gap is 3.2 eV), the rutile type TiO2 only can be irradiated by the light less than 413 nm wavelength, and excitation wavelength of anatase TiO2 response is shorter than 387.5 nm, which both belongs to the ultraviolet (UV) light region. The light utilization rate of solar light is only 5% [START_REF] Fang | Modifications on reduced titanium dioxide 123 photocatalysts: A review[END_REF] .

 OH +  OH →H2O2 (R2-6)
Secondly, the recombination rate of photogenerated carriers is very high, resulting in low light quantum efficiency. In addition, the photocatalyst is powdery, difficult to recover and reuse. It is also one of the important factors that restrict its development.

Therefore, in order to improve the energy utilization and quantum efficiency of TiO2, the problem of recovery has become a hot topic in the field of photocatalysis. As it is known from previous research that iron species are photosensitive and can enhance the cycling and transformation of reactive radicals in the environment, hence it could obviously be expected that iron species can be used in AOPs to participate in wastewater treatment. The processes including photochemical process employing Fe(III)/Fe(II) species, photocatalytic processes in the presence of iron oxides, Fenton and photo-Fenton processes have already been widely investigated and proved to be efficient to produce HO  . Among all these processes, Fenton and photo-Fenton processes are the most promising and attracting methods, since this kind of process is proved to be green, efficient, easy to operate and low-cost, leaving water and carbon dioxide as byproduct. The reaction between H2O2 and Fe(III)/Fe(II) is a well-known method for the generation of the hydroxyl radical (HO  ), including Fenton (Fe(II)/H2O2) [START_REF] Sherwood | Modified Fenton oxidation of diesel fuel in arctic soils rich in organic matter and iron[END_REF][START_REF] Huang | Combination of Fenton processes and biotreatment for wastewater treatment and soil remediation[END_REF] , Fenton-like (Fe(III)/H2O2) [START_REF] Huang | Effect of ethylenediamine-N,N'-disuccinic acid on Fenton and photo-Fenton processes using goethite as an iron source: optimization of parameters for bisphenol A degradation[END_REF] and photo-Fenton (UV/Fe(III)/H2O2) reactions [START_REF] Giannakis | Effect of Fe(II)/Fe(III) species, pH, irradiance and bacterial presence on viral inactivation in wastewater by the photo-Fenton process: Kinetic modeling and mechanistic interpretation[END_REF][START_REF] Huang | Development of a new homogenous photo-Fenton process using Fe(III)-EDDS complexes[END_REF][START_REF] Wu | Fe(III)-EDDS complex in Fenton and photo-Fenton processes: from the radical formation to the degradation of a target compound[END_REF][START_REF] Santos-Juanes | Combining ZVI reduction with photo-Fenton process for the removal of persistent pollutants[END_REF][START_REF] Wu | Photodegradation of 4-tert octylphenol in aqueous solution promoted by Fe(III)[END_REF] , the reaction equation and second order constant was shown below (R12-25). As it is known to all, Fenton process was found more than 100 years ago, and application to destroy the structure of toxic organic compounds was started from late 1960s. It is found that the Fenton reaction could be very efficient in the removal of many hazardous organic pollutants which were resistant to biological treatment from water in a short time and could completely remove the contaminants to harmless compounds such as CO2, H2O and inorganic salts. With high oxidation abilities, HO  (E 0 =2.8 V) has a wide application in removing environment contaminates like endocrine disrupter, chlorophenol, dye wastewater, heavy metals, pharmaceuticals, pesticide and landfill leachate due to its nonselectivity to organic chemicals. 58, 65, 99, 100 . 101 In recent decades, the mechanism of Fenton process has been widely used in inclusive number of wastewater treatments research such as (1) homogeneous Fenton and Fenton-like systems using different kinds of Fe(II) and/or Fe(III) salt and peroxide in acidic medium; (2) homogeneous photo-Fenton system by using UV irradiation to reduce Fe(III) to Fe(II) 102 ; (3) heterogeneous Fenton and Fenton-like system using clay, zero valent iron or iron oxides; (4) electro-Fenton or photo-electro-Fenton system. The effect of variables such as temperature, pH value, initial concentration and inorganic ions was also determined in these related researches 101,[103][104][105][106][107][108] .

Photo-catalyst + hv→h VB + +e CB - (R2-7) h VB + +e CB -→ energy(heat) (R2-8) h VB + +H 2 O/OH -→HO  +H + (R2-9) O 2 +e CB -→ O 2 •- (R2-10) HO  /O 2 •-/ h VB + +Pollutant → Intermediates→H 2 O+CO 2 (R2-11)
Fe 2+ + H2O2 → Fe 3+ + OH -+  OH k=76 M -1 s -1 (R2-12)
Fe 3+ + H2O2 → Fe 2+ + HO2  + H + k=0.02 M -1 s -  OH + HO2  → H2O + O2 (R2-23)

 OH + O2 -→OH -+ O2 (R2-24)  OH +  OH → H2O2 (R2-25) 2-3-3-2. SO4 •-radical
While in recent years, studies on sulfate radical (SO4 •-) have proved that it's very effective in degrading recalcitrant organic pollutants due to its higher oxidation-reduction potential (E 0 =2.6-3.2 V) [START_REF] Wu | Sulfate Radical Photogeneration Using Fe-EDDS: Influence of Critical Parameters and Naturally Occurring Scavengers[END_REF] , longer half-life period and more selective abilities than HO  . Although the research on SO4

SO 4 •-+OH -→HO  +SO 4 2- k=(6.5±10)×10 7 M -1 s -1 (R2-32) SO 4 •-+H 2 O→HO  +SO 4 2-+H + k=11.7 M -1 s -1 (R2-33) SO 4 •-+SO 4 •-→S 2 O 8 2- k=4×10 8 M -1 s -1 (R2-34) SO 4 •-+S 2 O 8 2-→S 2 O 8 •-+ SO 4 2- k=6.1×10 5 M -1 s -1 (R2-35) SO 4 •-+Fe(II)→SO 4 2-+Fe(III) k=4.6×10 9 M -1 s -1 (R2-36)

2-3-4. Degradation of pharmaceutical pollutant by AOPs

In recent years, advanced oxidation processes (AOPs) have been proposed as alternative methods to remove persistent organic pollutants (POPs) in different environmental media [113][114][115] . Generally, AOPs can achieve the recalcitrant pollutant destruction when biological treatments are unfeasible, such as for endocrine disrupting chemicals (EDCs) 116,117 , pharmaceuticals and personal care products (PPCPs) 118,119 . Main oxidation processes are usually based on the in situ production high reactive species such as hydroxyl (HO  ) and sulfate radical (SO4 •-) which are able to oxidize organic pollutants in aqueous solutions 113,114 . In the last two decades different AOPs based on iron activation under dark 120,121 , UV light 122 and/or in the presence of radical precursors 123 have been proposed demonstrating their efficiency for organic pollutants abatement. On the other hand, heterogeneous semiconductor photocatalysis is also one of the most promising oxidation process. Under irradiation, electron (e-) and hole (h+) pairs are produced in conduction and valence band, leading to the formation of reactive oxygen species (ROS) such as HO • or hydroperoxyde/superoxide radical anion (HO2 • /O2 •-, pka 4.88) 124 . Nowadays, the degradation of pharmaceuticals by AOPs included these methods as follows, and commonly used AOPs in degrading high detected pharmaceutical pollutants are listed in Table 23451) Photocatalytic degradation. In the presence of catalyst, organic compound in wastewater can be decomposed by strong oxidant species. Some organic compound can be directly mineralized into CO2 and H2O. Some refractory organics can also be converted into biodegradable molecules, or molecules can be direct photolyzed. This process can significantly decrease COD, while BOD/COD increase, increasing the biodegradability of wastewater, and making it suitable to be discharged. The treatment of pharmaceutical industry wastewater by catalytic oxidation can overcome the insignificant effect of traditional biochemical treatment of pharmaceutical wastewater, and effectively attack the conjugated bonds of organic molecules, so as to achieve the goal of removing COD and improving biodegradability 125 . In the photocatalytic degradation processes, the choice of catalyst is the most important step. So, in recent years, more and more researchers concentrate on the modification, enhancement and optimization of photocatalysts. It's well known that carbamazepine is one of the most difficult removal pharmaceutical in traditional microbiological technics such as STP or MBR [START_REF] Radjenovic | Analysis of pharmaceuticals in wastewater and removal using a membrane bioreactor[END_REF] . While, it can be removed effectively by photocatalytic methods. [START_REF] Ghasemian | Photoelectrocatalytic degradation of pharmaceutical carbamazepine using Sb-doped Sn80%-W20%-oxide electrodes[END_REF]126 Hongche et al. Results showed that, the efficiency of UV/H2O2 was higher than direct photolysis, and the remove efficiency can increase with the increased addition of H2O2. While under the same conditions, Photo-Fenton can increase the efficiency by 20% compared with Fenton. Nowadays, the synthesis of BiOX include solve-thermal, precipitation at room temperature 159 , sol-gel 160 and hydrolysis 161 . During the process of synthesis, various solvent are adopted such as water, ethanol, methanol, isopropanol, ethylene glycol, diglycol, glycerol 154,162,163 . In order to increase the surface area of catalyst and form oriented space structure, some surface active agent such as polyvinylpyrrolidone (PVP) 164 , polyacrylamide (PAM) 165 and cetyltrimethylammonium bromide (CATB) 166 .

The synthesized BiOX possessed great surface area, and pore volume, enhancing the

2-5. Fe(III)-Ethylenediamine-N,N'-disuccinic acid (EDDS)

2-5-1. Iron in water

Iron is a natural element present in different phases of the earth. In the Earth crust, iron is the second abundant metal following aluminum; iron also exists in living creature, and exhibits as one of the most important trace elements in human body which is essential for health; after all, iron is also one of the common compositions in natural aqueous phase, including rain, cloud water, fog, other atmospheric water and surface water. There were a lot of related reports about the iron detection in natural aqueous phase. In fact, iron is present under a variety of forms in water ranging from soluble to colloidal and particulate species. But, most of the iron in natural waters exists in the form of insoluble ferric oxides and (hydr)oxides. The concentration of dissolved iron is very low and most of the dissolved iron is associated with strong organic ligands in natural waters. It has been known that the iron species including Fe 3+ , Fe(OH) 2+ , Fe(OH)2 + and Fe2(OH)2 4+ /Fe2O 4+ exist in aqueous Fe(III) salt solution, however, actual species in natural water depend on different conditions. In the aqueous solutions at pH < 5, there are at least four kinds of Fe(III) complexes with different forms, including Fe 3+ , Fe(OH) 2+ , Fe(OH)2 + and Fe2(OH)2 4+ . The distribution of these complex forms is described by the following equilibria [START_REF] Brillas | Electro-Fenton Process and Related Electrochemical Technologies Based on Fenton's Reaction Chemistry[END_REF]206 :

Fe 3+ + H2O  Fe(OH) 2+ + H + (R2-37) Fe 3+ + 2H2O  Fe(OH)2 + + 2H + (R2-38) 2Fe 3+ + 2H2O  Fe2(OH)2 4+ + 2H + (R2-39)
Iron can also absorbs sun radiation and effect on the environmental chemistry through photoinduced process is of great relevance. The different products from hydrolysis of iron salts have different photochemical property. At pH around 3,

[Fe(H2O)5(OH)] 2+ is the predominant species, which exhibits significant photoactivity in the UV and visible region of solar radiation. It can generate HO  through photolysis, and the quantum yield of the reaction is determined to be 0.075 at 360 nm 206 , whereas the photochemical reduction could be shown as following (R2-40). And Fe(II), photoreduced from Fe(III), can be oxidized again to Fe(III). Except the most reactive Fe(OH) 2+ , other Fe(III)-OH complexes are also photosensitive.

[Fe(H2O)5(OH)] 2+ + h → Fe 2+ + HO  (R2-40)

2-5-2. Iron complex

As reported above, the generation of HO  and SO4 •-species using iron sources, mainly combined with UV irradiation, has played predominant roles in advanced oxidation processes. However, there are several key parameters in these activation methods. The most predominant drawback is the fact that Fe 3+ ions precipitate at pH higher than 4.0 and usually under these conditions Fe 2+ is naturally oxidized into Fe 3+ .

Therefore, most of the iron species presenting in natural and slightly basic solution, exist in the form of insoluble ferric oxides and (hydr)oxides [START_REF] Wu | Fe(III)-EDDS complex in Fenton and photo-Fenton processes: from the radical formation to the degradation of a target compound[END_REF]207 showed that the degradation efficiency of Fe (III)-oxalate/UV system was much higher. Hug et al. 212 studied the oxidation removal effect of various oxidants on arsenic trivalent As (III) under the neutral condition (pH=6.5-8), UVA ultraviolet was taken as radiation source. The results showed that As (III) could not be oxidized under the condition of only Fe (III) or H2O2. When citrate was added to the system, the removal efficiency of As (III) was greatly improved due to the formation of the complex of iron citrate (Fe (III)-citrate).

2-5-3. Ethylenediamine-N,N'-disuccinic acid (EDDS)

Very recently different research groups started to work with a new complexing agent of iron EDDS, its physical and chemical properties are shown in The use of the S,S-EDDS as chelant for Ca 2+ , Sn 2+ , Cu 2+ , Fe 3+ and Zn 2+ was widely investigated in previous literatures. In 2002, total production of EDDS achieved10000 tons. Nowadays, EDDS is a chelating reagent available in Europe on a large scale, and the demand for S,S-EDDS acid has been growing at the rate of 15% per year 214 .

[S, S]-EDDS [R, S]-EDDS or [S, R]-EDDS

HO 2 C H N H CO 2 H H N H CO 2 H CO 2 H [R, R]-EDDS
Our group has studied the physiochemical properties of the Fe(III)-EDDS complex with a ratio of 1:1. Under irradiation, Fe(III)-EDDS was easily photolyzed leading to the formation of HO  in the wide pH range between 3 and 9. As shown in Fig. 234567, the molecule of EDDS contains four carboxyl functional groups, consequently, EDDS ionization equilibrium in the solutions is described in following reactions (R41)-(R44). For simplification, EDDS molecules were represented as H4L 215,216 .

The physicochemical properties of the Fe(III)-EDDS complex are also investigated. When Fe (III) is complexed with EDDS, the generated Fe(III)-EDDS can be presents as four main forms FeL -, Fe(OH)L 2-, Fe(OH)2L are the same as for any other iron species. On the contrary, the quantum yield of HO  radical formation was higher at higher pHs between 3.0 and 9.0. This result is particularly interesting in terms of the natural environment.

H4L → H + +H3L - pKa1 = 2.4 (R2-41) pollutants and heavy metals in water and soil near neutral pH [START_REF] Huang | Development of a new homogenous photo-Fenton process using Fe(III)-EDDS complexes[END_REF][START_REF] Wu | Fe(III)-EDDS complex in Fenton and photo-Fenton processes: from the radical formation to the degradation of a target compound[END_REF]111,217,[219][220][221][222][223] . In the previous study of our group, Huang et al (2013)

H3L -→ H + +H2L 2- pKa2 = 3.9 (R2-42) H2L 2-→ H + +HL 3- pKa3 = 6.8 (R2-43) HL 3-→ H + +L 4- pKa4 = 9.8 (R2-44)

Characterization of Bismuth catalyst

The crystal phase structure was characterized by X-ray diffraction (XRD, Bruker

Photodegradation of ACTP

The irradiation setup was the same for p-HPA. The solutions were magnetically stirred with a magnetic bar during irradiation, and the total volume of the solution was 100 mL. All the experiments were carried out at room temperature (293 ± 2 K)

supported by a circulating cooling water system. Samples were taken from the reaction tube at fixed interval times and the catalyst solid was removed by 0.2 𝝁m PTFE filters before analysis. In the experiments, the concentration of ACTP was always stayed at 50 𝝁M, concentration of H2O2 and Na2S2O8 was always 1 mM and BiOCl0.75I0.25 catalyst was kept at 0.3 g L -1 .

Analytical and spectroscopic methods

p-HPA detection

The concentration of the p-HPA remaining in the aqueous solution was determined by an Alliance high performance liquid chromatography (HPLC) equipped with a photodiode array detector (Waters 2998, USA) and Waters 2695 separations module. The flow rate was 0.15 mL min -1 , the injection volume was 50 μL and the mobile phase was a mixture of water (0.1% H3PO4 included) and methanol (65/35, v/v). The column was a Nucleodur 100-3 C18 of 150×2.0 mm, particle size of 3 μm. Under these conditions p-HPA was detected at 274 nm and retention time was 6.7 min. Calibration of p-HPA was shown in Fig. 3-3. The concentration of BQ was determined by HPLC-UV using the same column and mobile phase was a mixture of water and methanol (75/25, v/v) with a detection wavelength set at 245 nm. The retention time of BQ is 4.9 min, and for calibration see Fig. 34.

UV-vis spectra were recorded with a Cary 300 UV-visible spectrophotometer.

pH values of the solutions were measured using a Cyberscan 510 pH meter.

Mineralization of p-HPA solution was determined by total organic carbon (TOC) analysis (5050A, Shimadzu, Japan). The limit of detection and limit of quantification have been determined to be 50 and 17 µgC L -1 , respectively, considering the signal-to-noise ratio. Fig. 34The standard curve of BQ

ACTP detection

The concentration of the ACTP remaining in the aqueous solution was determined by high performance liquid chromatography (HPLC) (Alliance, Waters, 2695, USA) equipped with a photodiode array detector (Waters 2998). A Nucleodur 100-3 C18 reverse column (150×2.0 mm, 3.0 𝝁m) was used to separate the compound in the solution. The flow rate was 0.15 mL min -1 with a mobile phase mixture of methanol and 0.1% H3PO4 contained water (30/70, v/v). The UV detection of ACTP was set at 244 nm wavelength, and the retention time was 4.9 min, and for calibration see Fig. 345.

UV-vis spectra were recorded with a Cary 300 UV-visible spectrophotometer. pH values of the solutions were measured using a Cyberscan 510 pH meter. Fig. 345The standard curve of ACTP

Laser Flash Photolysis

Experiments were carried out using the fourth harmonic (λexc = 266 nm) of a Quanta Ray GCR 130-01 Nd:YAG laser system instrument and the energy was set at 45 mJ/pulse. Other conditions are kept the same to those described in previous articles [START_REF] Wu | Sulfate Radical Photogeneration Using Fe-EDDS: Influence of Critical Parameters and Naturally Occurring Scavengers[END_REF]225 . Two different solution were prepared at pH 2.5 and 8.5 (milli-Q water adjusted with HClO4 and NaOH). High concentrated stock solutions (p-HPA, ACTP, Na2S2O8, H2O2 and SCN -) were mixed just before each experiment and diluted with Mili-Q water to obtain the desired mixtures and concentrations in a 3 mL quartz cuvette. For the detection of second order rate constant between p-HPA, ACTP and HO  , SCN - (0.1 mM), H2O2 (5.5 mM) and different concentrated p-HPA and ACTP were mixed in the quartz cell. The detection wavelength was set at 475 nm to detect maximum absorbance of SCN  . For second order rate constant between p-HPA and SO4 The catalyst was synthesized through a precipitation method in ambient pressure in EG-H2O mixed solvent 226 , BiOCl0.75I0.25 crystals showed flower-like morphologies.

Compared with BiOX catalyst researched before, BiOCl0.75I0.25 possessed greater surface area and total pore volume determined by N2 sorption 159,229 , facilitating the adsorption of p-HPA molecules on the surface and supplying abundant active cites for the oxidation ability. Introduction of iodine to BiOCl crystalline lattices induces the formation of heterojunction, and decreases the recombination rate of hole-electron pairs. Different to BiOCl alone, the band-gap of BiOCl0.75I0.25 was adjusted to 2.12 eV, which is suitable for the utilization of simulated solar light (λ > 300 nm).

Optical properties of p-HPA and blank experiment

The UV-vis spectra of p-HPA (500 µM) at different pH conditions are shown in Fig. 45, in which can be seen that p-HPA has a maximum absorb at 274 nm. When pH was lower than 9, the absorb edge located at 300 nm, indicating that it can't absorb visible light. While, when pH of the solution was changed to 10.

Reactive species generation

Reactive species generated under irradiation of BiOCl0.75I0.25 catalyst are determined using competition kinetic approach with specific scavengers in solution.

Isoproponal (IPA), potassium iodide (KI) and benzoquinone (BQ) are taken as HO • M -1 s -1 ) 238 and concentration adopted for the desorption experiment (2 mM), we can find that this reaction is few orders of magnitude lower than the reaction between HO • and p-HPA in the solution.

Effects of dissolved oxygen

Effects of dissolved oxygen on p-HPA degradation during the photocatalytic degradation process is investigated by supplying proper gas (oxygen and nitrogen) to the reaction solutions before (for 20 min) and during experiments. Despite no significant differences in the adsorption of pollutants on the catalyst, the degradation of p-HPA is accelerated with the increasing of oxygen concentration (Fig. 4-14). In fact, the pseudo-first order degradation constant is found to increase from (2.4 ± 0.5) × 10 -3 min -1 in N2 saturated to (2.4 ± 0.2) × 10 -2 min -1 in O2 saturated solutions. This trend indicates that oxygen plays a key role in the photocatalytic degradation process through the formation of HO2 • /O2 •-species.

Effects of Fe 3+ and Na2S2O8 addition

In traditional AOPs methods, Fe 3+ 98 and S2O8 were widely studied in the removal of organic pollutant through the generation of HO  and SO4 •-with high oxidation abilities. While research on the combination effects of BiOX with Fe 3+ and S2O8 2-was rare up to now. In this paper, the degradation of p-HPA by Fe 3+ alone and combined with BiOCl0.75I0.25 catalyst were investigated at pH 3.0 to avoid precipitation (Fig.4-15). Under irradiation, Fe 3+ can produce HO  through a series of chain reactions (R4-2), which would oxide p-HPA fast in the first 15 min followed with a very slow degradation because of the consumption of Fe 3+ . When Fe 3+ and BiOCl0.75I0.25 both existed in the solution, the adsorption was similar, while the degradation rate of p-HPA unexpectedly decreased compared with catalyst alone. The pseudo first order constant supplied in Table 4-3. To explain this problem, reactions between Fe 3+ and HO2 • /O2 •-radicals must be taken into consideration (R4-3 and R4-4) 122 . Although Fe 3+ donated the generation of HO  , which possessed higher oxidation potential than HO2 • /O2 •-, while in current conditions the quantum yield of HO  was limited and HO2 • /O2 •-was consumed a lot to the transformation between Fe 2+ and Fe 3+ , thus competing with p-HPA molecules and decreasing the removal efficiency.

Fe(OH) 2+ +hv→Fe(II)+HO  (R4-2) Fe(III)+HO 2 • →Fe(II)+O 2 +H + k<1.0×10 3 M -1 s -1 (R4-3) Fe(II)+HO 2 • +H 2 O→Fe(III)+H 2 O 2 +OH -k=1.2×10 6 M -1 s -1 (R4-4)
To explore the possible synergic effects on the p-HPA removal using the Na2S2O8/BiOCl0.75I0.25/UV system, UV/ Na2S2O8 efficiency was firstly investigated, as shown in Fig. 45678910111213141516. Approximately 30% of p-HPA can be removed in two hours when pH changed from 3.0 to 9.8, indicating activation of S2O8 2-by UV is not directly correlated to the pH. However, when BiOCl0.75I0.25 catalyst was added, the degradation efficiency increased compared with catalyst alone (see Table 4-3), especially in neutral and alkaline solutions. This result was totally different to the pH effects in Fig. 4567. The increased degradation can be ascribed to the following reasons.

Firstly, the direct photolysis of S2O8 2-generating SO4 •-(E 0 =2.6-3.2 V) contributed to the oxidation of p-HPA. Secondly, S2O8 2-also can trap the photogenerated conduction band electron results in the formation of SO4 These results also suggest that BiOCl0.75I0.25 catalyst is able to decompose the by-products of p-HPA to CO2 and H2O. After 2 hours of irradiation TOC removal percentage is around 45% and do not significantly change after 3 more hours, suggesting that some by-product can be adsorbed on the surface of catalyst, occupying the adsorption site. However, after 5 hours of irradiation, complete degradation of p-HPA is observed. HO2 • /O2 •-couple seems to be less reactive on the by-products than on p-HPA. p-HPA by-product generated during the photocatalytic process were detected by GC-MS after solvent extraction. 4-hydroxybenzyl alcohol and p-hydroxybenzaldehyde are identified as main intermediates (Table 4-4) on the basis of molecular mass and literature data [START_REF] Miranda | Pyrylium salt-photosensitized degradation of phenolic contaminants present in olive oil wastewater with solar light: Part III. Tyrosol and p-hydroxyphenylacetic acid[END_REF][START_REF] Minh | Degradation of olive oil mill effluents by catalytic wet air oxidation: 2-Oxidation of p-hydroxyphenylacetic and p-hydroxybenzoic acids over Pt and Ru supported catalysts[END_REF] .

Conclusion

In this work, BiOCl0.75I0.

Comparison of hydroxyl and sulfate radical degradation efficiencies of p-HPA activated by Fe(III)-EDDS/UV : influence of critical parameters

The reaction between H2O2 and Fe(III)/Fe(II) is a well-known method for the generation of hydroxyl radicals (HO  ), including Fenton (Fe(II)/H2O2) [START_REF] Sherwood | Modified Fenton oxidation of diesel fuel in arctic soils rich in organic matter and iron[END_REF][START_REF] Huang | Combination of Fenton processes and biotreatment for wastewater treatment and soil remediation[END_REF] , Fenton-like (Fe(III)/H2O2) [START_REF] Huang | Effect of ethylenediamine-N,N'-disuccinic acid on Fenton and photo-Fenton processes using goethite as an iron source: optimization of parameters for bisphenol A degradation[END_REF] and photo-Fenton (UV/Fe(III)/H2O2) reactions [START_REF] Giannakis | Effect of Fe(II)/Fe(III) species, pH, irradiance and bacterial presence on viral inactivation in wastewater by the photo-Fenton process: Kinetic modeling and mechanistic interpretation[END_REF][START_REF] Huang | Development of a new homogenous photo-Fenton process using Fe(III)-EDDS complexes[END_REF][START_REF] Wu | Fe(III)-EDDS complex in Fenton and photo-Fenton processes: from the radical formation to the degradation of a target compound[END_REF][START_REF] Santos-Juanes | Combining ZVI reduction with photo-Fenton process for the removal of persistent pollutants[END_REF][START_REF] Wu | Photodegradation of 4-tert octylphenol in aqueous solution promoted by Fe(III)[END_REF] .

However, in recent years, studies on sulfate radical (SO4 -) have proved that this radical is outstanding in degrading recalcitrant organic pollutants ascribing to its similar oxidation-reduction potential (E 0 = 2.6-3.2 V) [START_REF] Wu | Sulfate Radical Photogeneration Using Fe-EDDS: Influence of Critical Parameters and Naturally Occurring Scavengers[END_REF] , longer half-life time and more selective reactivity compared with HO  . From previous literatures, the generation of SO4 -usually derives from the activation of persulfate (PS) by heat [START_REF] Nie | Degradation of chloramphenicol by thermally activated persulfate in aqueous solution[END_REF] , UV or transition metals [START_REF] Nie | Degradation of chloramphenicol by persulfate activated by Fe2+ and zerovalent iron[END_REF]109,110 . The generation of HO  and SO4 -based on iron source, mainly combined with UV irradiation, has played predominant roles in advanced oxidation processes. So the development of Fe(II)/Fe(III) organic complexes is essential to improve the removal efficiencies in higher pH solutions. Photo-production of HO  and SO4 -based on Fe(II)/Fe(III)-EDDS complexes has been widely studied in recent years, showing excellent efficiency in removing organic pollutants and heavy metals in water and soil near neutral pH.

Our laboratory has studied the physicochemical properties of the Fe(III)-EDDS complex. Fe(III) is complexed by EDDS with a ratio 1:1. Results showed that the use of Fe(III)-EDDS complexes leads to efficient oxidations whatever the sources of radical species (HO • or SO4 -). However, the comparison between Fe(III)-EDDS/H2O2/UV and Fe(III)-EDDS/PS/UV systems under same conditions has never been studied before, including the effects of pH, Fe(III)-EDDS concentration, H2O2 and PS concentrations. The main goal of this study is to compare these two processes using p-HPA as a model pollutant. Kinetics degradation of p-HPA were investigated in detail with Fe(III)-EDDS/H2O2/UV and Fe(III)-EDDS/PS/UV systems. The aim of this study is to quantify the different activation efficiency of H2O2 and PS by Fe(III)-EDDS under irradiation in the same experimental conditions and to correlate with the production of HO  and SO4 - radicals. Especially to reach this goal and understand well the different mechanisms.

Photochemistry of Fe(III)-EDDS

As it is known from previous research of our group the best stoichiometry between Fe(III) and EDDS in Fe(III)-EDDS complex is 1:1 217 , the stock solution of Fe(III)-EDDS (5 mM) was prepared by mixing 10 mM EDDS solution with 10 mM Fe 3+ solution together. It is observed that Fe(III)-EDDS complex in water solution is stable for at least 10 days when kept in 4 ℃. The UV-visible spectrum of Fe(III)-EDDS (200 uM, pH 3.9) complex solution stayed at different days showed in (molecular form) and pH 8.5 (anionic form), are adopted to determine the second order rate constant. Under 266 nm laser excitation of p-HPA alone, there is no detected transition species in the solution between 300-600 nm (∆ 20 nm). The determination of the second order rate constant between p-HPA and HO  (k p-HPA,HO•) is determined by using chemical competition kinetics with thiocyanate anions (SCN -) and di-thyociante radical anions (SCN2 -) specie, as shown in equations R5-1-R5-4

and Scheme 5-1. k p-HPA,HO• is obtained by following the absorbance of SCN  at 475 nm with different concentration of p-HPA added. The absorbance decreases when p-HPA concentration increases because p-HPA compete with SCN -for the reaction of HO  . The slope "a" can be obtained from the linear regression of Abs0/Abs versus different p-HPA concentrations and thus the second order rate constant kp-HPA,HO•under two different pH's 2.5 and 8.5 can be evaluated (Table 5-1). Details concerning the second order rate constant determinations are reported in Fig. 5-2. kp-HPA•SO4•-(M -1 s -1 ) (4.77±0.09)×10 9 (3.51±0.12)×10 9

H 2 O 2 +hν→2HO • (R5-1) HO • +SCN -→HO -+SCN • (R5-2) HO • +p-HPA→product (R5-3) SCN • +SCN -→SCN 2 •- (R5-4)
Scheme 5-1: second order constant calculation process, where Abs is the absorbance at 475 nm with different concentrated p-HPA, Abs0 is the absorbance in the absence of p-HPA. kHO•,,SCN-=1.2×10 10 M -1 s -1 241 . [P] is the concentration of p-HPA.

To obtain kp-HPA•SO4•-, transient absorption spectra of SO4 •-obtained are recorded after 266 nm laser excitation of S2O8 2-in the absence or presence of p-HPA.

Maximum absorption of SO4 •-occurred at 450 nm, and the first-order decay constants of the radical were determined as it is described in our previous paper [START_REF] Wu | Sulfate Radical Photogeneration Using Fe-EDDS: Influence of Critical Parameters and Naturally Occurring Scavengers[END_REF] . At pH 2.5, a decrease of the transient species is observed with a pseudo-first order constant kSO4•-:4.2910 4 s -1 . After the addition of p-HPA (3.3310 -4 M), the transient decay increases to 1.58×10 6 s -1 , indicating that SO4 •-and p-HPA react with each other with high yield. The second order rate constant kp-HPA•SO4•-can be calculated using a linear regression of the pseudo-first order decay constant of SO4

•-(k', s -1 )
versus the p-HPA concentration (Fig. 5-2B). The same method is applied at pH 8.5

and the results are gathered in Table 5-1. It is clear that pH can affect kp-HPA,HO• and kp-HPA,SO4•, showing faster reaction activities in acidic solution (molecular form of p-HPA) than alkaline solution (anionic form of p-HPA). This effect is more pronounced with HO  , kp-HPA,HO• is 3 times higher at pH 2.5 than at pH 8.5, than with SO4 •-, kp-HPA•SO4•-is only around 1.3 times higher at pH 2.5 than at 8.5. ,

x P Abs k a k SCN                                • • - - H O ,P H O ,SC N k = a k [SC N ]

Effect of UV and Fe(III)-EDDS complex

In the dark, p-HPA is stable in aqueous solution even in the presence of Fe(III)-EDDS solution. Moreover, p-HPA is not degraded under UV, because it cannot absorb the wavelengths longer than 300 nm, no photolysis process is observed in our experimental conditions. As a contrary, under UV irradiation used in this study, H2O2

and Na2S2O8 can produce HO  (R5-1) and SO4 •-(R5-5).

S 2 O 8 2-+ hν→2 SO 4 •- (R5-5) Fe 2+ +H 2 O 2 →Fe 3+ +HO  +OH -k = 76 M -1 s -1 (R5-6) Fe 3+ +H 2 O 2 →Fe 2+ +HO 2 • +H + k = 0.02 M -1 s -1 (R5-7) S 2 O 8 2-+Fe 2+ →SO4 •-+SO 4 2-+Fe 3+ k = 20-27 M -1 s -1 (R5-8) Fe(III)-EDDS+H 2 O 2 →Fe(II)-EDDS+ HO 2 • O 2 •- ⁄ +H + (R5-9) Fe(III)-EDDS+hν→Fe(II)+EDDS • (R5-10)
However, p-HPA has a relatively low degradation rate in the presence of H2O2 ( 1mM) or Na2S2O8 (1 mM) alone under UV irradiation, achieving 10% and 30%

remove respectively, after 120 min of irradiation (Fig. 5-3 and 5-4). In fact, in our irradiation conditions (300 to 500 nm), the quantum yield of HO• formation is very low and so the kinetic of p-HPA degradation. In terms of pH effect, faster p-HPA degradation is observed at pH 3.0 or 4.5 than at pH 9.4 in agreement with the evaluated second order rate constants as a function of pH (Table 5-1).

Fig. 5-5A shows the degradation of p-HPA with Fe(III)-EDDS (100 µM) at two different H2O2 concentrations (100 and 500 µM) with and without UV. The disappearance of p-HPA is much faster in Photo-Fenton (Fe(III)-EDDS/H2O2/UV) system than in Fenton-like process (R5-7, kFe(III),H2O2 = 0.02 M -1 s -1 ) 111 . According to the Fenton reaction (R5-6), the production of HO  is correlated with the formation of Fe(II) (R5-9 and R5-10) [START_REF] Wu | Fe(III)-EDDS complex in Fenton and photo-Fenton processes: from the radical formation to the degradation of a target compound[END_REF]111 and so with the efficiency of pollutant oxidation. Under irradiation, UV increases and contributes a lot to the generation of Fe(II), inducing the fast decomposition of p-HPA in the first 10 min through the Fenton process (R5-6) [START_REF] Huang | Development of a new homogenous photo-Fenton process using Fe(III)-EDDS complexes[END_REF][START_REF] Wu | Fe(III)-EDDS complex in Fenton and photo-Fenton processes: from the radical formation to the degradation of a target compound[END_REF] . At higher H2O2 concentration (500 µM) added into the solution, p-HPA can be removed completely in 60 min in Fenton-like process and in 10 min in photo-Fenton system. At lower concentration of H2O2 (100 µM) p-HPA is not removed completely of the solution with or without UV irradiation. These results indicate that H2O2 concentration is a crucial and a limiting parameter in the oxidation process.

For the experiments with Na2S2O8, pH and the different concentrations are kept the same compared with systems using H2O2. The different degradation kinetics are shown in Fig. 5-5B. Although some previous papers have reported that Fe(III) can activate persulfate to produce SO4 •-in the dark, it was not the case in our system. In fact, much higher concentration of Fe(III) species and persulfate seem needed to do this activation 207 . Under irradiation, Fe(III)-EDDS/Na2S2O8 system shows a good efficiency for p-HPA degradation but much lower compared to the system with H2O2.

Different reasons can explain the lower efficiency in the presence of Na2S2O8. First of all, it is important to mention that the formation of Fe(II) via reaction R5-10 is one of the crucial step to active S2O8 2-or H2O2 to produce oxidative species through reactions R5-6 and R5-8, which are identified and evaluated in terms of relative importance in the following part of this paper. The lower efficiency can be due to the fact that the second order rate constant of the activation of S2O8 2-(kNa2S2O8•Fe(II)= 20-27 M -1 s -1 ) is more than three times smaller than the one with H2O2 (kH2O2•Fe(II)=76 M -1 s -1 ). In addition, the lower efficiency of Na2S2O8 can be also due to the second order rate constants between p-HPA with HO  or SO4 •-, kp-HPA,HO• is four times higher than kp-HPA,SO4• in acidic solutions (Table 5-1). As a contrary, the photolysis of H2O2 and Na2S2O8 were not responsible of this effect. Indeed, with the concentration and light irradiation used this process of photolysis is negligible and moreover is more efficient with Na2S2O8 than H2O2. [Fe(III)-EDDS] = 100 𝝁M , pH = 3.9.

Effects of H2O2 and Na2S2O8 concentrations

Initial degradation rate of p-HPA (Rp-HPA) is adopted to evaluate the degradation efficiency of the reaction process. Rp-HPA is evaluated from the first 5 min of irradiation, because during this period pH of the solution kept stable, and the degradation can be well fitted by pseudo first order kinetics. Fig. 5-6A shows the initial degradation rate of p-HPA with different H2O2 concentrations in photo-Fenton process. Rp-HPA increases from 1.5010 -7 to 3.8010 -7 M s -1 when H2O2 concentrations increases from 50 µM to 1 mM. The observed increase of p-HPA degradation when the concentration of H2O2 increased is not a surprise considering that H2O2 is one of the two compounds generating HO  in the Fenton process and as described in Fig. 5-5A, H2O2 concentration was the limiting parameter. However, H2O2 is also a scavenger of HO  (kH2O2•HO•= 2.710 7 M -1 s -1 ) and so too high concentration of H2O2 will lead to a decrease of the organic compound degradation rate [START_REF] Huang | Development of a new homogenous photo-Fenton process using Fe(III)-EDDS complexes[END_REF] . Due to the much higher rate constant of HO  reaction on p-HPA (kp-HPA•HO• = 2.210 10 M -1 s -1 ) the decrease of the p-HPA degradation is not observed in our experimental condition; at our highest H2O2 concentration used (1 mM) 90% of HO  is still consumed by p-HPA.

So it was only observed that the increase rate slowed down when H2O2 increased from 250 µM to 1 mM. Moreover, at this concentration the photolysis of H2O2 is still negligible (Fig. 5-3) and do not strongly interact in the p-HPA degradation.

The effect of Na2S2O8 concentration is shown in Fig. 5-6B. Similar to H2O2, the degradation rate is always increasing when Na2S2O8 concentration increases. In fact S2O8 2-is the source of radical species and the negative effect of Na2S2O8, in terms of radical species scavenger probably SO4 •-in this case (kNa2S2O82-,SO4•-=6.1×10 5 M -1 s -1 ) 109 , is much slower than in the case with H2O2 (kH2O2•HO•= 2.710 7 M -1 s -1 ). So in our experimental conditions ([Na2S2O8 2-]  1 mM) a constant increase of the degradation rate of p-HPA with the increase of persulfate concentration is observed. of Fe(II) followed with higher concentration of HO • through Fenton process. However, EDDS is one kind of organic matter and is able to interact in the process. Compared with H2O2, EDDS has an almost 100 times higher second order rate constant with HO • (kEDDS,HO•=2.48±0.43×10 9 M -1 s -1 >> kH2O2,HO•=2.7×10 7 M -1 s -1 ) 242 . So, EDDS and its photo-induced by-products can compete more strongly with p-HPA molecules.

Moreover, another reason can explain the inhibition at higher Fe(III)-EDDS concentration. In fact, Fe(III)-EDDS is decomposed very fast under irradiation 217,218 , and the high concentration of Fe(II) produced is oxidized by different ways and the formed Fe(III) species are spontaneously precipitated at the used pH (7.5). Thus the formed colloid or precipitation can be responsible for the inhibition of the penetration of photons in the solution.

In Fe(III)-EDDS/Na2S2O8/UV system, Na2S2O8 is set as 500 µM, because at lower concentration (100 µM), it is difficult to differentiate the effect of Fe(III)-EDDS concentration due to relatively low degradation percentages. From [Na2S2O8] = 500 𝝁M, pH = 3.9.

Effects of pH

In the Fe(III) (photo)-induced activation reactions, pH always plays a significant role during the whole process. Firstly, in the p-HPA/Fe(III)-EDDS/UV system, pH effects are investigated, as shown in Fig. 5678. Under irradiation ( > 300 nm), Fe(III)-EDDS can produce HO  through a series of reactions as it is described in Abida et al. (2006) 243 and as it is proved in different studies like in the paper of Li et al. (2010) 218 . However, if we compare with photo-Fenton process, the photo-production of HO  is very limited without H2O2 and so the removal of p-HPA is much slower (Fig. 5-5)

When pH increases from 3.0 to 4.7, the removal efficiency increases very fast, while with pH further increase up to 9.4, degradation percentage continues to increase but much more slowly. The constant increase of the percentage of p-HPA removal with the increase of pH is due to the higher HO  formation quantum yield when the pH increases as it is demonstrated in the paper of Li et al. (2010) 218 . The slower increase observed at pH higher than 4.7, can be attributed to the lower reactivity (three times lower) of HO  on the mono-anionic form of p-HPA. The rate constants have been evaluated for the first time during this study and are presented in Table 5- So, the activation mechanisms of these two different processes are not exactly the same.

In Fe(III)-EDDS/H2O2/UV reaction, p-HPA degradation efficiency increases fast from pH 2.5 to 7.5, then it begin to decrease when pH increases to 9.0. As explained with the system Fe(III)-EDDS/UV (Fig. 5-8), in general, the increase of efficiency with the increase of pH is mainly due to the increased HO  formation quantum yield.

However, in the photo-Fenton process HO  generation is mainly due to the Fenton process and so to the photo-generation of Fe(II) which is correlated with the first photochemical step, the photoredox process from the complex Fe(III)-EDDS leading to the formation EDDS  and Fe(II). To compare and understand well this mechanism, Fe 3+ is introduced instead of Fe(III)-EDDS, see Fig. 5-10A. In order to keep Fe 3+ soluble in the solution, a very acidic pH is selected (pH = 2.0). Fe(III)-EDDS shows faster degradation efficiency than Fe 3+ in the initial stage, ascribing to the higher photochemical activity of the complex Fe(III)-EDDS than the ion Fe 3+ . Like for Fe(III)-EDDS the degradation is also very fast in the first 10 minutes and after slows down. In Fe 3+ /H2O2/UV system, the degradation nearly stopped after about 30 minutes of irradiation. This important slow down and then complete stop of the reaction is due to the almost total consumption of H2O2. The higher final removal percentage observed with Fe 3+ than with Fe(III)-EDDS is due to the presence of EDDS which can consume also HO• and so decrease the reactivity efficiency on p-HPA.

As a contrary of this rapid initial degradation observed with H2O2, in the Fe(III)-EDDS/Na2S2O8/UV system, a gradual degradation during the irradiation process is observed (Fig. 5-5). Fe(III)-EDDS is decomposed fast almost in the first 10 min by photoredox process, into Fe(II) and EDDS  ,whatever the species present in the solution [START_REF] Wu | Sulfate Radical Photogeneration Using Fe-EDDS: Influence of Critical Parameters and Naturally Occurring Scavengers[END_REF]217 . In the Fe(III)-EDDS/Na2S2O8/UV system Rp-HPA slightly increases until pH 4.0 and then decreases when the pH increases (Fig 5-9B). The pH effect is not exactly the same than with H2O2 (increase is observed until around pH 8.0). This difference of pH effect could be due to the interaction of S2O8 2-in the radical chemistry processes. Indeed, S2O8 2-can react with the first radical generated by the photoredox process on EDDS (R5-12) and then avoid the reaction with O2 leading to the formation of HO2

Identification of active species at pH 3.9

In order to determine the radicals involved in the degradation of p-HPA in the different processes, IPA and TBA are used as scavengers of radicals and so added in the solution (Fig. 5-11). In fact, at adopted concentrations, IPA (10 or 20 mM) was considered to quench efficiently both generated SO4 •-and HO  , considering the second order rate constants of kIPA,SO4•-= 7.42 × 10 7 M -1 s -1 and kIPA,HO• = 1.9 × 10 9 M -1 s -1 , while TBA (1 or 2 mM) can be considered to be more selective toward HO  (kTBA,HO• = 6.0 × 10 8 M -1 s -1 ) than with SO4 •-(kTBA,SO4•-= 8.31 × 10 5 M -1 s -1 ) 207 .

From In the presence of persulfate (500 µM), 20 mM of IPA can react with 86% SO4 •- when p-HPA is present at 50 µM, so higher concentrated IPA is needed if SO4 •-want to be completely inhibited. HO  radical is mainly generated from two routes. Firstly, it is generated from the photolysis of Fe(III)-EDDS complex. Secondly, SO4 •-can also react with hydroxyl anion or water molecule to produce HO  (R15-16) 109,112 . 245 To evaluate the relative significance of the two radicals (sulfate and hydroxyl) in this system, experiments are also performed with TBA acting specifically as a trap for hydroxyl radical. When the concentration of TBA increases from 1 to 2 mM the decrease of p-HPA concentration near 60% is the same (Fig. 5-11B). So, from From In general, the p-HPA degradation efficiency is much higher with Fe(III)-EDDS/H2O2/UV system than with Fe(III)-EDDS/Na2S2O8/UV system whatever the solution pH's. Firstly, this difference can be explain by taking into account the second order rate constants kp-HPA,HO• and kp-HPA,SO4•-into consideration (Table5-1). The value of kp-HPA,HO• is higher than kp-HPA,SO4• both at pH 2.5 (4.6 times) and 8.5 (2 times). Secondly the rate constants of the key reactions generating the radical species is more than 3 times higher for the Fenton process (R5-6) than for the activation of persulfate with Fe(II) (R5-8). However, at pH 2.5, the ratio of the rate constants between radical species and p-HPA kp-HPA,HO•/ kp-HPA,SO4•-equal to 4.6 and the ratio of p-HPA disappearance rate in the two systems Rp-HPA (H2O2)/ Rp-HPA (Na2S2O8) equal to 5.2 are similar. It is slightly higher for the p-HPA disappearance rate which is coherent with the two reasons mentioned at the beginning of this paragraph. As a contrary the difference between these two ratios is much higher at pH 8.5, Rp-HPA (H2O2)/ Rp-HPA (Na2S2O8) equal to 16.7 is eight times higher than kp-HPA,HO•/ kp-HPA,SO4•-equal to 2.1. So, as we demonstrated in the paper of Huang et al. 111 , the Fenton process, involving Fe(III)-EDDS complex, is higher near neutral pH than in acidic pH. This significant result in terms of environmental aquatic compartments seems not present for the activation of persulfate to generate SO4 •- radical. Indeed, in this case a dramatic decrease of the efficiency is observed. So, in this particular study, with p-HPA used as organic pollutant and UV/Fe(III)EDDS as source of Fe(II), we clearly demonstrated that the Fenton process is more efficient than activation of persulfate process and more particularly at environmentally closed pH values. As one of the most important pain killer mainly sold as over the counter drugs, acetaminophen (ACTP) is heavily used all over the world, ranking as one of the top three drugs in England 246 . Different researches based on advanced oxidation processes such as UV/H2O2 58 , Fenton [START_REF] De Luna | Kinetics of acetaminophen degradation by Fenton oxidation in a fluidized-bed reactor[END_REF] , electron Fenton 59 and electrochemical degradation 115,247,248 of ACTP are present in literature. Furthermore, heterogeneous photocatalytic degradation efficiency and mechanism of ACTP by TiO2 was also widely used 246,249,250 . In TiO2 photocatalytic system, HO  is always the predominant active species 249 .

Degradation of

As described in Chapter 4, BiOCl0,75I0.25 possessed the greatest surface area and most degradation will be explained in detail. Furthermore, Second order constant between ACTP with HO  and SO4 •-radicals was determined by laser flash photolysis (LFP) or chemical competition reactions.

Blank experiment of ACTP mixed with H2O2 and Na2S2O8 in the dark

To explore ACTP degradation with H2O2 and Na2S2O8 under UV irradiation, blank experiments must be conducted in the dark firstly to check its stability, as shown in

ACTP degradation in UV/H2O2 and UV/Na2S2O8 systems under different pH values

The time evolution of Acetaminophen concentration in the presence of H2O2 and Na2S2O8 was shown in Fig. 6-2, and degradation profiles followed pseudo first order reaction kinetics. Firstly, direct photolysis of ACTP could be neglected (Fig. 6-2a).

UV-vis spectrum of ACTP at different pH was shown in Fig. 6-3, in which can be seen that, at pH 3 and 6.5, it showed same spectrum with the highest adsorption at 244 nm.

While when increased the solution pH to 11.2 a red shift with a maximum at 256 nm can be observed. In our photochemica reactor in which the emission spectrum is between 300 and 500 nm, ACTP has almost no adsortion. With H2O2 and Na2S2O8 addition, ACTP showed different degradation kinetics under various pH. Results of quenching experiments demonstrated that HO  and SO4 •-were the premoninant radicals involved in the UV/H2O2 and UV/Na2S2O8 system (Eq.1-2) at pH 5 respectively by taking IPA and EtOH as HO  (kIPA, HO•-=1.9× 10 9 M -1 s -1 ) 207 and SO4 •- (kEtOH, SO4•-=1.6-7.7× 10 7 M -1 s -1 ) [START_REF] Tan | Kinetic study of acetaminophen degradation by UV-based advanced oxidation processes[END_REF] scavengers respectively. In UV/H2O2 system, at neutral pH the maximum removal percentage is oberved. As contrast, in UV/Na2S2O8 system, degradation rate increased with increasing of pH (Fig. 6-2b). At least two reasons were involved simulaneously to account for this phenomenon. As reported in Fig. 6-3, ACTP under anion form was easier to be oxided. Furthermore, in alkaline solution the generation of HO  from SO4 •-is favoured (Eq.3-4) 109 . So, alkaline solution can provide a suitable environment to support the oxidation of ACTP by HO  and SO4 •-radicals, as in the UV/Na2S2O8 system. However, at very high pH values (>11) hydrogen peroxide (pKa = 11.7) is presnt as hydroperoxide anion (HO2 -). This form also showed like a scavenger of hydroxyl radicals and cause decomposition of H2O2. It was reported that the reaction of hydroxyl radicals with hydroperoxide anions is approximately 100 times faster than its reaction with H2O2 252 . In addition, too basic solution make H2O2 unstable and accelerate its self decomposition 56 . (Eq.5-7)

Furthermore, it has been reported that the standard redox potential of HO  in acidic pH is about 2.4-2.7 eV which is higher than that in alkaline pH (1.9-2.0 eV) 253 .

However, persulfate radical redox potential does not depend on the pH 245 . Effects of halide ions were widely studied in previous articles, especially chloride and bromide [START_REF] Li | The roles of halides in the acetaminophen degradation by UV/H2O2 treatment: Kinetics, mechanisms, and products analysis[END_REF]254 . While comparision of Cl -and I -, which were commonly detected in costal seawater, saline water or indsutrial wastewater discharges, were investigated for the first time by taking different concentrations in UV/H2O2 and UV/Na2S2O8 systems, respectively. Firstly, ACTP concentration kept unchanged when NaCl, H2O2 or Na2S2O8 are mixed under dark conditions (see Fig. 6-4). Under irradiation, the effects of Cl -are shown in Table 6-1, in which can be seen that in UV/H2O2 system, when 10 mM of NaCl are added, degradation rate slighly increased, while no differences can be observed using lower NaCl concentration was used in agreement with results obtained by Li et al. [START_REF] Li | The roles of halides in the acetaminophen degradation by UV/H2O2 treatment: Kinetics, mechanisms, and products analysis[END_REF] . In UV/Na2S2O8 system, Cl -always exhibted negligible effects at various concentrations. In previous works, Cl -plays a role of radical scavengers leading to the formation of chloride reactive species (Cl•,Cl2• -or ClHO• -) in the presence of HO  or SO4 •-255 56, 82, 252 .

H 2 O 2 +hν→2HO • (R6-1) S 2 O 8 2-+ hν→2SO4 •- (R6-2) SO4 •-+ OH -→HO • + SO 4 2-k=(6.5±1.0)×10 7 M -1 s -1 (R6-3) SO4 •-+ H 2 O→HO • +HSO 4 - k=11.92 M -1 s -1 (R6-4) H 2 O 2 ↔HO 2 -+H + pKa=11.7 (R6-5) HO 2 -+HO • →HO 2 • +OH - (R6-6) 2H 2 O 2 →2H 2 O+O 2 (R6-7) 0
When iodide ions were added to the solution, degradation rate was significantly affected as shown in Fig. 6-5. In the dark, the addition of KI (260 µM-5 mM) accelerate the degradation of ACTP in the presence of H2O2 or Na2S2O8 at pH 5.4, and H2O2 showed higher oxidation rates than Na2S2O8. Similar to HOCl, HOI was supposed to be genertaed from iodine ions by H2O2 and Na2S2O8 (Eq.8-11). It's well known that HOI is a very strong and unstable oxidants, and it can react with ACTP molecules 252 . Under As shown in Fig. 6-6, under UV irradiation, no degrdation of ACTP was observed in the presence of KI. However, when H2O2 or Na2S2O8 were added into the solution at same time, the degradation was highly enhanced. Except for the existence of HOI, reactive halogenated species can be generated from HO  or SO4 •-as shown in Eq.12-14 256 . Different from chloride radicals, iodide induced radicals (I•, HOI• and I2 •-) showed higher degradation reactivities with organic molecules. Tamtam et al. 254 have demonstrated that reactive halogenated species reacted more selectively than hydroxyl radicals with electron-rich organic pollutants in agreement with Li et al. [START_REF] Li | The roles of halides in the acetaminophen degradation by UV/H2O2 treatment: Kinetics, mechanisms, and products analysis[END_REF] . As acetaminophen contains hydroxyl and acetamido groups, which are electron donors to the aromatic rings, it's hypothesized in this research that these elecron-rich moieties may selectively react with these iodide related radical reactive species. Using much higher concentrations of iodide ions a decrease of rate was observed due to the self quenching effects as reported in Eq.14. To determie the second order constant between ACTP and HO  at pH 5.4, laser flash photolysis was adopted and the method was shown in section 3.6. Under 266 nm laser excitation of ACTP alone, there is no detected transition species in the solution between 300-600 nm (∆ 20 nm). The determination of k ACTP,HO• was based on the Eq.15-18, and obtained by following the absorbance of SCN  at 475 nm with different concentration of ACTP added. The absorbance decreased from 0.021 to 0.006 when ACTP concentration increased from 0 to 2.33×10 -4 M, because ACTP compete with SCN -for the reaction of HO  . The slope "a" can be obtained from the linear regression of Abs0/Abs versus different ACTP concentrations (see Fig. 678910111213), and it was 11276. Thus the second order constant kACTP,HO• can be evaluated from the product of "a", concentration of SCN -(1×10 -4 M) and kHO•,,SCN-=1.2×10 10 241 , finally it was calculated to be 1.35×10 A chemical competition method was adopted to determine the second order constant between ACTP and sulfate radical, and bisphenol A (BPA) was used as the reference compound (see Fig. 6-14), kSO4•-,BPA=1.5×10 9 M -1 s -1 has been determined in previous studies [START_REF] Sanchez-Polo | Comparative study of the photodegradation of bisphenol A by HO(•), SO4(•-) and CO3(•-)/HCO3 radicals in aqueous phase[END_REF]257,258 . For this experiment, the initial concentration of BPA was 50 uM in the presence of 50 uM ACTP and 1 mM Na2S2O8 under UV irradiation at pH

I -+H 2 O 2 /S 2 O 8 2-→I 2 +H 2 O/SO 4 2- (R6-8) I 2 +H 2 O↔HOI+I -+H + (R6-9) I 2 +H 2 O 2 →HOI (R6-10) 3HOI→2I -+IO 3 -+3H + (R6-11) HO•+I -→HOI •- k=1.1×10 10 M -1 s -1 (R6-12) I•+I -→I 2 •- k=8.8×10 9 M -1 s -1 (R6-13) I•+I•→I 2 k=1.5×10 10 M -1 s -1 (R6-14) 6

Conclusion

In this work, different homogeneous and heterogeneous AOPs were adopted to the degradation of ACTP, which were all based on the UV irradiation. Firstly, ACTP was stable alone in solution, while it showed a rapid oxidation when mixed with H2O2 and Na2S2O8 when pH exceeded 10, indicating easily reacts with ionic form. Faster degradation rate was observed at pH 5.4 in UV/H2O2 and at pH 9.5 in UV/Na2S2O8.

The addition of chloride ions (10 mM) had negligible effects on these two systems, while low concentrated iodide ions can accelerate the degradation. This trend was attributed to the fact that reactive halogenated species reacted more selectively than hydroxyl radicals with electron-rich organic pollutants such as ACTP. The addition of Fe(III)-EDDS greatly increased the degradation rate in a short time due to the fast activation between Fe(II) with H2O2 and Na2S2O8. ACTP showed fastest removal rate at pH 7.5 in UV/H2O2/Fe(III)-EDDS and at pH 3.2 in UV/Na2S2O8/Fe(III)-EDDS, which can be attributed to the photochemical properties of Fe(III)-EDDS under different pH. Secondly, in the heterogeneous photocatalytic degradation of ACTP, BiOCl0.75I0.25 was taken as a semiconductor catalyst we synthesized before. ACTP showed the fastest removal at pH 3, because acidic solution favored the transformation of electrons to the surface of catalyst, followed with closed to the surface. kACTP,HO• was determined by LFP to be 1.35×10 10 M -1 s -1 , and kSO4•-,ACTP was determined to be (1.94±0.1)×10 9 M -1 s -1 when competing with BPA.

Conclusion

In In Fe(III)-EDDS/H2O2/UV reaction, p-HPA degradation efficiency increased fast from pH 2.5 to 7.5, then it began to decrease when pH increased to 9.0. While Rp-HPA began to decrease when pH increased to 3.9 in Fe(III)-EDDS/Na2S2O8/UV. HO  was the main active species in photo-Fenton process, and both SO4 •-and HO  were responsible for the degradation in Fe(III)-EDDS/Na2S2O8/UV system. In general, Fe(III)-EDDS/H2O2/UV always showed higher degradation efficiency than ACTP had the fastest removal rate at pH 3, because acidic solution favored the transformation of electrons to the surface of catalyst, followed with the formation of HO2 • /O2 •-, which was the main active species prevailed. Interfacial mechanism research indicated that the degradation occurred mainly on the surface of catalyst or in the bulk solution that closed to the surface. kACTP,HO• was determined by LFP to be 1.35×10 10 M -1 s -1 , and kSO4•-,ACTP was determined to be (1.94±0.1)×10

  et al investigated the photocatalytic degradation of p-HPA by pyrylium salt. At pH 3.0 and the addition of 5% pyrylium salt,the remove percentage can achieve 50% in 350 min. The researchers also studied the degradation of p-HPA by ozone and UV. Irama Sanche et al. 51 studied the catalytic degradation of p-HPA by zero valance iron, and the degradation efficiency was checked by the detection of COD and TOC of the wastewater. Minh, D. Pham et al. 52 investigated the degradation of p-HPA and p-hydroxybenzoic acid, which were two most commonly detected pollutants in olive oil industrial wastewater. The authors studied the catalytic oxidation in a batch reactor using platinum and ruthenium catalysts supported on titanium and zirconium oxides at 140 ℃ and 50 bar of total air pressure.

Fig. 2 - 2 Fig. 2 - 3

 2223 Fig.2-2 Structure of p-hydroxyphenylacetic acid (p-HPA)

Figure 2

 2 Figure 2-4 basic mechanism of TiO2 photocatalytic process

2 )

 2 photolysis, this method utilizes oxidations O3 or H2O2 activated by UV or visible light with the assistant of Fe 2+ , Fe 3+ or iron complex, to finally produce HO•. Hilla et al.130 investigated the degradation of metronidazole by UV, Fenton and Photo-Fenton.

(

  photocatalytic performance. The above two characteristics can effectively reduce the recombination rate of photoelectron hole pair of bismuth oxide 150 . Bismuth halide is a highly anisotropic layered structure semiconductor. The crystal structure of bismuth halide is PbFCl type, which belongs to the tetragonal system. X -and O 2-distribution around Bi 3+ , the two inverse square columns coordination, Cl-layer is Affirmative coordination, the next layer is O 2-layer, Bi 3+ layer sandwiched in the middle, the crystal structure of BiOCl as shown in Figure2-6 143 , the three-dimensional lattice model, each ion independent occupy a level. As "the five layer Cl-Bi-O-Bi-Cl structure", with van Edward, the large molecules tend to laminated structure. The crystal structure of BiOX can also be seen as the double X -ion layer and Bi2O2 layer alternately arranged along the c axis. The halogen atoms are bonded by halide atoms through non chemical bond force, and the binding force is relatively small, so it is easy to dissociate.

Figure 2 - 6

 26 Figure 2-6 Structure diagram of BiOCl (small ball-Bi 3+ , middle ball-O 2-, big ball-Cl -)

  photocatalytic abilities than the 2D materials. Solvents evidently performed important functions as structure-directing agents in the catalyst synthesis process because of their special physical and chemical properties. The majority of solvents involved in the reaction media included certain types of alcohols, especially polyols like ethylene glycol, diethylene glycol, and glycerol.The production of reactive species under light excitation are also different, including the photo-induced holes, hydroxyl and superoxide radicals. The use of semiconductors on the degradation of pollutants mainly relates to dyes, pharmaceutical wastewater, phenols, and heavy metals and so on. Synthesis and degradation of BiOX material are summarized as shown in Table2-6.

Fig. 2 - 7

 27 Fig. 2-7 Chemical structures of three different stereoisomers of EDDSCurrently, S,S-EDDS is used as laundry detergent (the tri-sodium salt) to replace phosphorous containing chelant, in hair products to stabilize peroxide solutions and as enhancer of biocides, for the metal removal in wastewater and soil washing treatment.

2 - 8 .

 28 photoactive species in natural water. For the first time, the quantum yield of HO  was detected by photolysis of Fe(III)-EDDS under UV irradiation. The quantum yield of HO  was independent of the concentration of Fe(III)-EDDS. The effects of oxygen and irradiation wavelength on the quantum yield of HO  were very well known and

Figure. 2 - 8 2 - 5 - 4 .

 28254 Figure. 2-8 The proposed theoretical model distribution of the predominant species for the Fe(III)-EDDS complex as a function of pH

  demonstrated that in homogeneous Fenton process (Fe(III)-EDDS/H2O2), Fe(III)-EDDS (molar ration 1:1) showed better removal efficiency of bisphenol A (BPA) in alkaline solution than in acidic one111 . At the same time,Wu et al (2014) demonstrated that under UV, photo-Fenton process (Fe(III)-EDDS/H2O2/UV) has much higher efficiency in removing 4-tert-Butylphenol (4-tBP) than Fenton process. This result can be explained with the fast generation of Fe(II) under irradiation[START_REF] Wu | Fe(III)-EDDS complex in Fenton and photo-Fenton processes: from the radical formation to the degradation of a target compound[END_REF] . In the presence of PS with Fe(III)-EDDS/UV, efficient oxidation of 4-terbutylphenol (4-tBP) was observed due to the formation of SO4 •-. In neutral and basic pH conditions, the efficiency of 4tBP degradation was much higher with Fe(III)-EDDS than with Fe(III) aqua complexes[START_REF] Wu | Sulfate Radical Photogeneration Using Fe-EDDS: Influence of Critical Parameters and Naturally Occurring Scavengers[END_REF] . So the use of Fe(III)-EDDS complexes leads to efficient oxidations whatever the sources of radical species (HO  or SO4 •-).Miralles-Cuevas et al.[START_REF] Miralles-Cuevas | Pharmaceuticals removal from natural water by nanofiltration combined with advanced tertiary treatments (solar photo-Fenton, photo-Fenton-like Fe(III)-EDDS complex and ozonation)[END_REF] demonstrated that mild solar photo-Fenton combined with EDDS (low starting concentration of iron, no pH modification and low consumption of H2O2), could be a good option for the tertiary treatment of nan filtration concentrates from municipal wastewater treatment plants containing low concentrations of pharmaceuticals. Photo-Fenton-like Fe(III)-EDDS complex makes it possible to work at pH over 6.Papoutsakis et al. 223 verified that the treatment of micro-contaminants with the use of EDDS as a n iron-complexing agent is possible in real wastewater effluent at a pH range between 5 and 8. Yan et al.209 investigated the effectiveness and mechanisms of naphthalene and metal removal from artificially contaminated soil by FeEDTA/FeEDDS-activated persulfate through batch experiments. Using FeEDTA-activated persulfate, higher naphthalene removal from the soil at 7 hours was achieved 89%, compared with Fe(III)EDDS-activated persulfate (75%). Although EDDS is advantageous over EDTA in terms of biodegradability, it is not preferable for iron chelate-activated persulfate oxidation since persulfate was consumed to oxidize EDDS.Zhang et al. 219 assessed the application of calcium peroxide activated with Fe(II)-EDDS to enhance trichloroethylene degradation in aqueous solution. And the results indicated that EDDS prevented soluble iron from precipitation, and the optimum molar ratio of Fe(II)/EDDS to accelerate pollutant degradation was 1:1. deionized water, maintaining their total molar amount equal to the molar amount of Bi(NO3)3.5H2O. The aqueous solution was added dropwise into the previous EG solution, and the pH of this mixture was adjusted to 9.0 with NH3. Thereafter, the solution was continually stirred at 80 °C for 3 hours. Finally, the mixture was separated by centrifugation, washed several times with deionized water and ethanol, dried at 60 °C for 12 hours. To improve the repeatability of the experiment, samples were calcined at 200 °C for 2 h to remove the residual EG. The synthesised materials could be regarded as BiOCl0.75I0.25.

  D8 Advance) using Cu Kα radiation(40 kV and 40 mA) as X-ray source (λ= 0.154056 nm) with a 2θ scope of 10-90°. The morphology of the photocatalyst was observed through field emission scanning electron microscopy (FESEM,Ultra 55, Germany). Surface area and pore distribution of the catalysts was determined by N2 adsorption and desorption, using the multipoint Brunauer-Emmett-Teller (BET) analysis methods by an AUTOSORB-1 N2 adsorption apparatus (Quantachrome, USA). The UV-vis diffuse reflection spectra (DRS) were obtained for the samples using a Scan UV-vis spectrophotometer (Thermo Nicolet Evolution 500 UV-vis) equipped with an integrating sphere assembly, and BaSO4 was used as reflectance sample. X-ray photoelectron spectroscopy (XPS) spectra are recorded with a PerkinElmer PHI 5000C ESCA system equipped with a hemispherical electron energy analyzer. The Mg Kɑ (hν = 1253.6 eV) anode was operated at 14 kV and 20 mA. The carbonaceous C 1s line (284.6 eV) was used as the reference to calibrate the binding energies (BEs).

1 . 3 - 1 .

 131 Photocatalytic degradation experiments were performed in a homemade photoreactor placed in a cylindrical stainless steel container. Four fluorescent bulb lamps (Philips TL D15W/05) were separately placed in the four different axes while the photoreactor, a water-jacketed Pyrex tube of 2.8 cm internal diameter, was placed in the center of the setup, as shown in Fig.3-1. Irradiation with the predominant wavelength between 300 to 500 nm was used to simulate solar light, and the emission spectrum is shown in Fig.3-2. The emission spectrum (FigureS1) reaching the solution was determined using an optical fiber coupled with a CCD spectrophotometer (Ocean Optics USD 2000+UV-VIS) calibrated using a DH-2000-CAL Deuterium Tungsten Halogen Reference Lamp. Energy has been normalized to actinometry results by using paranitroanisole/pyridine actinometer 224 .At the wavelength range of 300-500 nm, a total flux of 1451 W m -2 reaching the solution was determined.Solutions were magnetically stirred with a magnetic bar during irradiation, and the total volume of the solution was 100 mL. p-HPA adsorption on bismuth catalyst was investigated in the dark with continuous magnetic stirring. All the experiments were carried out at room temperature (293 ± 2 K) using a circulating cooling water system. An aliquot of solution (1 mL) was withdrawn from the reactor at fixed interval times and the catalyst solid was removed by 0.2 μm PTFE filters before analysis. Prior to the use of PTFE filter, control experiments were performed, showing that no adsorption of p-HPA was observed on these filters. Experiments are performed using a p-HPA concentration of 50 μM and BiOCl0.75I0.25 catalyst of 0.3 g L -1 . The effect of dissolved oxygen on the p-HPA degradation and radical species involvement are carried out bubbling N2 or O2 for 20 min before catalyst addition and during all the time of the experiment (with and without light).

Fig. 3 - 1 Fig. 3 - 2 3 . 4 . 2 . 3 - 1 .

 313234231 Fig. 3-1 Home-made photoreactor with four tubes

ForFig. 3 - 3

 33 Fig. 3-3 The standard curve of p-HPA

Fig. 4 - 1 4 - 2 .

 4142 Fig.4-1 shows the XRD patterns of the as-prepared BiOCl0.75I0.25 samples synthesized in EG-H2O media. All the samples can be indexed to a tetragonal phase BiOCl (JCPDS card NO. 06-0249), diffraction peaks of BiOI can be found hardly because of its relatively low content, and no other impurities can be observed. It can be deduced that solid solution has been formed in this composite catalyst because the similar diffraction peaks between BiOCl and BiOI, and it's not just a mixture between these two crystals.159, 226

Fig. 4 - 1 Fig. 4 - 2 4 - 3 . 4 -

 4142434 Fig.4-1 XRD patterns of BiOCl0.75I0.25 samples synthesized in EG-H2O media

Fig. 4 - 3

 43 Fig.4-3 XPS spectra of BiOCl0.75I0.25 sample: (A) survey spectra, (B) C 1s, (C) Bi 4f, (D) O 1s, (E) Cl 2p and (F) I 3d.

Fig. 4 -

 4 Fig. 4-4. (a) UV-vis DRS of the as-prepared BiOCl0.75I0.25 sample, (b) the band gaps (Eg) of BiOCl0.75I0.25 samples.

Fig. 4 - 5 Fig. 4 - 6 4 - 8 .Fig. 4 - 7 :Fig. 4 - 8 4 - 2

 454648474842 Fig.4-5 The UV-vis spectra of p-HPA (500 µM) at different pH conditions

Fig. 4 -Fig. 4 -

 44 Fig.4-9 p-HPA degradation with and without IPA. Conditions: [p-HPA] = 50 μM, [BiOCl0.75I0.25] = 0.3 g L -1 , [IPA] = 1 mM and pH = 3.0.

Fig. 4 -Fig. 4 - 4 . 5 .

 4445 Fig.4-11: p-HPA concentration decrease with (○) and without (•) BQ addition. Adsorption experiments were performed in the dark. Conditions: [p-HPA] = 50 μM, [BiOCl0.75I0.25] = 0.3 g L -1 , [BQ] = 2 mM and pH = 3.0.

Figure 4 -

 4 Figure 4-13: (A) Effect of NaH2PO4 addition on the p-HPA adsorption (first 60 min in the dark) and on p-HPA concentration in solution after photocatalytic experiment. (B)

Figure 4 - 14 :

 414 Figure 4-14: Effects of dissolved oxygen in the photocatalytic degradation of p-HPA by BiOCl0.75I0.25. Conditions were: [p-HPA] = 50 μM, [BiOCl0.75I0.25] = 0.3 g L -1 , pH = 3.0

Fig. 4 -Fig. 4 -Table 4 - 3 8 .

 44438 Fig.4-15 The degradation of p-HPA in the UV/Fe 3+ / BiOCl0.75I0.25 system, Conditions: [p-HPA]= 50 𝝁M, [BiOCl0.75I0.25] = 0.3 g L -1 , pH = 3.0, [Fe 3+ ]= 1mM

  25 has been prepared by precipitation method in EG-H2O mixed solvent. The as-prepared samples showed 3D flower-like morphology, possessing great surface area 34.93 m 2 g -1 and narrow band-gap 2.12 eV. Furthermore, adsorption and photocatalytic degradation of p-HPA by BiOCl0.75I0.25 in water have been investigated. Interfacial mechanism experiments reveal that photocatalytic degradation of p-HPA occurred mainly on the catalyst surface. pH strongly affects the adsorption percentage and degradation rate through changing the surface charge of catalyst and molecular form of p-HPA. p-HPA showed the maximum adsorption at pH 4.5 (Natural pH) which can be ascribed to the electrostatic attraction occurred between the positively charged p-HPA and negatively charged BiOCl0.75I0.25. While, p-HPA exhibited the faster degradation rate at pH 3. Dissolved oxygen concentration is a key parameter necessary to generate HO2 • /O2 •-, the main active species responsible for the pollutant degradation, which was identified by quenching experiment. Na2S2O8 showed synergistic effect with BiOCl0.75I0.25 under UV irradiation by supplying SO4 •-radicals. BiOCl0.75I0.25 shows promising p-HPA removal abilities but some photo-generated by-products still adsorbed on the surface of catalyst, seems not degraded. This photocatalyst, generating mainly HO2 • /O2 •-, is OH HO OH O very promising and can have strong implication in the treatment of polluted media in which hydroxyl radical should be strongly scavenged by organic matter or inorganic constituents.

Fig. 5 - 1 .Fig. 5 - 1 . 5 . 2

 515152 Fig.5-1. It showed a particular absorb at about 240 nm.

Figure 5 - 2 :

 52 Figure 5-2: (A) Ratio of absorption of transient at 475 nm in the absence (Abs0) and presence (Abs) of p-HPA used for determination of second order rate constant between hydroxyl radical and p-HPA. (B) Pseudo-first order rate constant of SO4 •- decay as function of different concentrated p-HPA.

Fig. 5 - 3 Fig. 5 - 4 Fig. 5 -

 53545 Fig.5-3 Degradation of p-HPA (50 𝝁M) in the UV/H2O2 (1 mM) system

Fig. 5 -Fig. 5 -

 55 Fig.5-6 (A) The H2O2 concentration effect on the initial degradation rate of p-HPA in photo-Fenton system at pH = 7.5, (B) The Na2S2O8 concentration effect on the initial degradation rate of p-HPA in Fe(III)-EDDS/Na2S2O8/UV system at pH = 3.9 . [p-HPA] = 50 𝝁M, [Fe(III)-EDDS] = 100 𝝁M, irradiation time considered to evaluate the degradation rate = 5 min.

Fig. 5 -FeFig. 5 -

 55 Fig.5-7B, a similar trend can be observed than with H2O2, so EDDS and its by-product also function as scavengers of SO4 •-radical (kEDDS, SO4•-=6.21× 10 9

1 .

 1 The pH effect on the Fe(III)-EDDS/H2O2/UV and Fe(III)-EDDS/Na2S2O8/UV processes during the initial stage (5 min of irradiation) are shown in Fig.5-9. The trend of pH effects is different from Fe(III)-EDDS/UV system and moreover the trend in Fig.5-9A and Fig.5-9B are also quite different although the experimental conditions are similar but the oxidation species used (H2O2 and Na2S2O8) are different.

Fig. 5 - 8

 58 Fig.5-8 Removal percentage of p-HPA in Fe(III)-EDDS/UV system after 120 min of irradiation at different pH. [Fe(III)-EDDS] = 100 𝝁M, [p-HPA] = 50 𝝁M.

Fig. 5 -Fig. 5 -

 55 Fig.5-9 (A) pH effect on the initial degradation rate of p-HPA in photo-Fenton system and (B) in Fe(III)-EDDS/Na2S2O8/UV system. [p-HPA] = 50 𝝁M, [Fe(III)-EDDS] = 100 𝝁M , [H2O2] = 100 𝝁M, [Na2S2O8] = 100 𝝁M, irradiation time considered to evaluate the degradation rate = 5min.

Fig. 5 -

 5 11A, 20 mM of IPA can almost completely inhibit the degradation of p-HPA. At this concentration, 97 % of HO  is consumed by IPA and only 3% was consumed by p-HPA. At lower concentration of IPA (1 mM) only 63% of HO  can react with IPA and so a degradation of p-HPA was still observed and correspond to almost 30% of removal. So, HO  radical is identified as the active species responsible of p-HPA transformation in the system with H2O2.

Fig. 5 -Fig. 5 -

 55 Fig.5-11B, it is possible to conclude both radicals are involved in the system Fe(III)-EDDS/Na2S2O8/UV with a specific percentage respectively near 20% for HO  and 80% for SO4 •-.SO4 •-+OH -→SO 4 2-+HO  k = (6.5±1.0)×10 7 M -1 s -1 (R5-15) SO4 •-+H 2 O→SO 4 2-+HO  +H + k = 11.92 M -1 s -1 (R5-16)

  suitable band-gap, showing excellent adsorption and photocatalytic degradation abilities in removing ACTP from water under simulated solar and visible light. In this system, O2 •-was the main radical determined by indirect scavenging experiments. In this work, three different oxidation species HO  , SO4 •-and HO2 • /O2 •-produced by UV/H2O2, UV/Na2S2O8, UV/H2O2/Fe(III)-EDDS, UV/Na2S2O8/Fe(III)-EDDS and UV/BiOCl0,75I0.25 respectively were studied. Under simulated solar irradiation (300 nm<𝝀<500 nm) their efficiency on the degradation of ACTP was compared for the first time. ACTP degradation in UV/H2O2, UV/Na2S2O8 and UV/BiOCl0,75I0.25 under different pH was studied, and influence of halide ions (Cl -and I -) were investigated at the same time. Interfacial mechanism, mineralization and efficiency of this

Fig. 6 - 1 .

 61 Fig.6-1. From Fig.6-1A, it can be seen that ACTP solution was stable between pH 3.0 and 9.5 within 40 hours. Moreover, when ACTP was mixed with H2O2 and Na2S2O8 (Fig.6-1B and Fig.6-1C), no significant degradation was observed when the nintial pH of the solution was lower than 10. At pH 11.5 a degradation in the first 2 hours ws observed (the pKa of ACTP was 9.38[START_REF] Tan | Kinetic study of acetaminophen degradation by UV-based advanced oxidation processes[END_REF] ). It was well known that H2O2 (E 0 =1.77 V) and Na2S2O8 (E 0 =2.01 V) are two common oxidants. So it can be deduced that ACTP was easily to be oxided in the ionized form. Su et al.[START_REF] Su | Degradation of acetaminophen in an aerated Fenton reactor[END_REF] evaluated ACTP removal in Fenton process, H2O2 (50 mM) alone has neglegible effects on ACTP degradation at pH 3. Zhang et al. 251 also studied the ACTP degradation in the presence of persulfate anion, and results showed that 35% of degradation was achieved with the initial concetration of PS set at 0.6 g L -1 at pH 6.5 within 40 min. In this study, control experiments on ACTP degradation was highly related to the solution pH.

Fig. 6 -

 6 Fig.6-1 (A) 50 𝝁M ACTP concentration as a function of time at different pH in the dark, (B) with Na2S2O8 (1 mM), (C) with H2O2 (1 mM).

Fig. 6 - 2 Fig. 6 - 3

 6263 Fig.6-2 Direct photolysis and degradations of ACTP (50 𝝁M) in the UV/H2O2 (1 mM) (A) UV/ Na2S2O8 (1 mM) (B) systems under different pH.

Fig. 6 - 2 , 6 - 7 .

 6267 only UV/H2O2 and UV/Na2S2O8 showed a low degradation rate of ACTP, especially in UV/H2O2 where about 10% of ACTP was removed in 2 hours. Fe(III)-EDDS is a potential iron complex in activating H2O2 and Na2S2O8, as explained in Chapter 5. So Fe(III)-EDDS mediated degradation of ACTP was discussed in this section, as shown in Fig.With a lower concentration of H2O2 and Na2S2O8 (500 µM), ACTP showed a fast degradation in these two systems and was completely removed in 15 min in UV/H2O2/Fe(III)-EDDS while 80% was removed in 30 min in UV/Na2S2O8/Fe(III)-EDDS. Compared to the direct photolysis of H2O2 and Na2S2O8, the addition of iron enhances the formation of HO  and SO4 •-, followed with a series oxidation reaction. With p-HPA, H2O2 showed faster degradation rate than Na2S2O8 because of the faster activation rate by Fe(II).

Fig. 6 - 7 9 6. 5 .Fig. 6 - 8 6 - 9 .Fig

 67956869 Fig.6-7 Degradation of ACTP in the UV/H2O2/Fe(III)-EDDS and UV/Na2S2O8/Fe(III)-EDDS Conditions: [ACTP] = 50 𝝁M, [Fe(III)-EDDS] = 100 𝝁M ,[H2O2]=[Na2S2O8]=500 𝝁M, pH = 3.9

Fig. 6 - 8

 68 Fig.6-8 pH effect on the initial degradation rate of ACTP in photo-Fenton system [ACTP] = 50 𝝁M, [Fe(III)-EDDS] = 100 𝝁M , [H2O2] = 100 𝝁M, irradiation time considered to evaluate the degradation rate = 5min.

Fig. 6 - 9 Fig. 6 -

 696 Fig.6-9 pH effect on the initial degradation rate of ACTP in Fe(III)-EDDS/Na2S2O8/UV system. [ACTP] = 50 𝝁M, [Fe(III)-EDDS] = 100 𝝁M , [Na2S2O8] = 500 𝝁M, irradiation time considered to evaluate the degradation rate = 5min.

Fig. 6 - 6 . 8 .

 668 Fig.6-12 The effects of adsorption and photocatalytic degradation of ACTP after the addition of NaH2PO4•H2O (a) and pseudo first order kinetics (b) Conditions: [ACTP]= 50 𝝁M, [NaH2PO4•H2O]= 2 mM, [catalyst]= 0.3 g/L, pH=3

5. 4 9 M

 49 for 240 min. At this pH conditions, sulfate radical was the predominant radicals, and there was neglible hydroxyl radical formation 112 . Then the concentration of BPA and ACTP were obtained at different treatment time intervials and the degradation rate constants of ACTP with sulfate radical were calculted using the above degradation rate constant of BPA and the following equation. Finally, kSO4•-,ACTP=(1.94±0.1)×10

Fig. 6 -Fig. 6 -

 66 Fig.6-13 Linear regression of Abs0/Abs versus different ACTP concentrations

  my PhD work, different homogenous and heterogeneous AOPs efficiency were investigated using representative pharmaceutical pollutants: ACTP and p-HPA. The interfacial mechanism was proposed in the BiOCl0.75I0.25 based photocatalytic degradation. In the Fe(III)-EDDS activated photo-Fenton and PS reactions, the different degradation efficiencies were evaluated under same conditions. Firstly, BiOCl0.75I0.25 has been prepared by precipitation method in EG-H2O mixed solvent. The as-prepared samples showed 3D flower-like morphology, possessing great surface area 34.93 m 2 g -1 and narrow band-gap 2.12 eV. Furthermore, adsorption and photocatalytic degradation of p-HPA by BiOCl0.75I0.25 in water have been investigated. Interfacial mechanism experiments reveal that photocatalytic degradation of p-HPA occurred mainly on the catalyst surface. pH strongly affects the adsorption percentage and degradation rate through changing the surface charge of catalyst and molecular form of p-HPA. p-HPA showed the maximum adsorption at pH 4.5 (Natural pH) which can be ascribed to the electrostatic attraction occurred between the positively charged p-HPA and negatively charged BiOCl0.75I0.25. While, p-HPA exhibited the fastest degradation rate at pH 3. Dissolved oxygen concentration is determined to be a key parameter necessary to generate HO2 • /O2 •-, the main active species responsible for the pollutant degradation, which was identified by quenching experiment. Na2S2O8 showed synergistic effect with BiOCl0.75I0.25 under UV irradiation by supplying SO4 •-radicals. BiOCl0.75I0.25 shows promising p-HPA removal abilities but some photo-generated by-products still adsorbed on the surface of catalyst, seems not degraded. This photocatalyst, generating mainly HO2 • /O2 •-, is very promising and can have strong implication in the treatment of polluted media in which hydroxyl radical should be strongly scavenged by organic matter or inorganic constituents. Furthermore, the problem of recycle use needs to be solved in the future.Secondly, the different activation efficiency of H2O2 and persulfate (PS) by Fe(III)-EDDS under irradiation to produce HO  and SO4 •-under same conditions were investigated in detail. Firstly, second order constant kp-HPA,HO• and kp-HPA,SO4• were determined by LFP at different pH. Result showed that both kp-HPA,HO• and kp-HPA,SO4• are higher at pH 2.5 than pH 8.5. Photo-Fenton (Fe(III)-EDDS/H2O2/UV) reaction was much faster than Fenton-like (Fe(III)-EDDS/H2O2, kFe(III),H2O2=0.02 M -1 s -1 ) ascribing to the fast formation of Fe(II) under irradiation. Under same conditions, Fe(III)-EDDS possessed lower activating efficiency of Na2S2O8 compared with H2O2 which can be due to the fact that the kNa2S2O8•Fe(II)= 20-27 M -1 s -1 is more than three times smaller than kH2O2•Fe(II)=76 M -1 s -1 . In addition, the lower efficiency of Na2S2O8 can also be due to the second order rate constants between p-HPA with HO  or SO4 •-, kp-HPA,HO• was higher than kp-HPA,SO4• at both acidic and alkaline solutions. The initial degradation rate of p-HPA (Rp-HPA) increased from 1.49×10 -7 M s -1 to 3.76×10 -7 M s -1 when H2O2 concentrations were ranging from 50 𝝁M to 1 mM, indicating that H2O2 dosage was the limiting parameter. The effect of Na2S2O8 concentrations was similar to H2O2, the degradation rate was always increasing, and the increase rate seemed unchanged between these concentrations. Using the same H2O2 and Na2S2O8 concentrations, the removal percentage of p-HPA increased when Fe(III)-EDDS concentration increased from 50 to 250 𝝁M. Then it decreased when further increased Fe(III)-EDDS concentration to 1 mM due to the scavenging effects of EDDS. The pH influence on the Fe(III)-EDDS/H2O2/UV and Fe(III)-EDDS/Na2S2O8/UV processes during the initial stage are evidently different.

Fe

  (III)-EDDS/Na2S2O8/UV in different pH solutions, especially in neutral and basic solution. Fe(III)-EDDS was a promising iron complex in practical wastewater treatment in natural conditions. Finally, different homogeneous and heterogeneous AOPs based on the UV irradiation were investigated to archive the degradation of ACTP. Firstly, blank experiments in dark showed that ACTP was stable alone in solution, while it can be rapidly oxidized when mixed with H2O2 and Na2S2O8 when pH exceeded 10, indicating that it was easily destroyed in ionic form. Under UV irradiation, it showed fastest degradation rate at pH 5.4 in UV/H2O2 and at pH 9.5 in UV/Na2S2O8. The addition of chloride ions (10 mM) had negligible effects on these two systems. While low concentrated iodide ions can accelerate the degradation efficiencies, which can be ascribed to the fact that reactive halogenated species reacted more selectively than hydroxyl radicals with electron-rich organic pollutants like ACTP. The addition of Fe(III)-EDDS greatly increased the degradation rate in a short time due to the fast activation between Fe(II) with H2O2 and Na2S2O8. ACTP showed fastest removal rate at pH 7.5 in UV/H2O2/Fe(III)-EDDS and at pH 3.2 in UV/Na2S2O8/Fe(III)-EDDS, which can be attributed to the special photochemical properties of Fe(III)-EDDS under different pH. Secondly, in the heterogeneous photocatalytic degradation of ACTP, BiOCl0.75I0.25 was taken as a semiconductor catalyst we synthesized before.

  [START_REF] Phan | Antibiotic contamination and occurrence of antibiotic-resistant bacteria in aquatic environments of northern Vietnam[END_REF]

Table 2 - 1

 21 Concentration of pharmaceuticals found in soil in different countries

		Range of concentration	Number of
	Country/region	(µg kg -1 )	samples	references
	Carbamazepine
	Mexico	2.6-7.5	4	Gibson et al
	Hebei,China	0.02-0.06	18	Chen et al 2011 12
	USA(a)	n.d a	3	Kinney et al
	USA(b)	0.7-1.4	1	Wu et al 2010
	Trimethoprim
	USA(a)	n.d-0.64	3	Kinney et al
	Hebei,China	0.64-2.15	18	Chen et al 2011 12
	Malaysia	3.1-60.1	10	Ho et al 2012 15
	Ibuprofen
	Mexico	n.d-0.1	4	Gibson et al
	Hebei,China	1.51-5.03	18	Chen et al 2011 12
	Diclofenac
	Mexico	n.d	4	Gibson et al
	Hebei, China	0.35-1.16	18	Chen et al 2011 12
	Sulfadiazine
	Hebei,China	1.15-3.82	18	Chen et al 2011 12
	Malaysia	n.d	10	Ho et al 2012 15
	Triclosan
	USA(b)	n.d	1	Wu et al 2010
	Mexico	n.d-16.7	4	Gibson et al
	a: undetectable

Table 2 - 2

 22 Occurrence and concentration of pharmaceuticals found in effluent, freshwater and groundwater from America, Europe and Asia.

					Range of concentration (ng L -1 )
			America			Europe			Asia
	Source	Effluen t WWTP /STP	Freshw ater-riv ers, canals	Ground water	Effluen t WWTP /STP	Freshwat er-rivers, canals	Groun dwater	Effluen t WWTP /STP	Freshwat er-rivers, canals	Groundw ater
	Location	New Mexico , Balkan	New Jersey, Canada	Califor nia, Canda	Spain, Italy	Span	Pan-E uropea n	Taiwan, Korea	Taiwan, Vietnarm
	Compound
	Antibiotics
	Trimethoprim	2550	145	18	39	59.9		2000	1808
	Ciprofloxacin		77	0.28	499	n.d a		2050		40
	Sulfamethoxazole	310	170	458	185	33	38	397	4330	3
	Analgesics
	Naproxen	1550	555			80.5
	Ibuprofen	11900	203	3.97	8000	468	395	1660	30
	Acetaminophen			1890		40
	Ketoprofen		16		940	293	2886	128	620
	Diclofenac	4200			740	794	24	1760	62
	Salicylic acid				3170	184
	Antiepileptic
	Carbamazepine	1550	735	420	150	366	3600	21000	120
	Stimulants
	Caffeine			290	300	568	4500	3180
		Brown
		et al,
	Reference	2006 18 Terzic
		et al,
		2008 7

Table 2 - 3

 23 Physical-chemical properties of p-hydroxyphenylacetic acid and

		acetaminophen
	English Name	p-hydroxyphenylacetic acid
	Abbreviation	p-HPA
	CAS number	156-38-7
	Molecular formula	C8H8O3
	Molecular weight	152.15
	Melting point(℃)	149-150
	pKa	4.5, 10.5
	Solubility	Slightly soluble in water
	English Name	acetaminophen
	Abbreviation	ACTP
	CASnumber	103-90-2
	Molecular formula	C8H9NO2
	Molecular weight	151.16
	Solubility(20 ℃)	14 g/L
	LogKow	0.46
	pKa	9.38

Table 2 -4

 2 Standard Reduction Potentials in Aqueous Medium of the Most

		Commonly Reported Oxidizing Agents 71
	NO.	Oxidizing agent	Standard reduction potential(E 0 /V)
	1	F2	3.05
	2	HO 	2.80
	3	SO4 •-	2.60-3.20
	4	Cl•	2.40
	5	FeO4 2-	2.20
	6	O3	2.08
	7	S2O8 2-	2.01
	8	H2O2/H2O	1.76
	9	H2O2/OH-	0.88
		MnO4 -(I)	1.67
		HO2 •-(I)	1.65
		HCO3 • or CO3 •-	1.59-1.63
		MnO4 -(II)	1.51
		HOCl	1.48
		HO2 •-(II)	1.65-1.44
		Cr2O7 2-	1.36
		Cl2 •-	1.36
		Cl2	1.36
		MnO2	1.23
		O2	1.23
		Br2	1.07
	2-3

-2. AOPs based on heterogeneous semiconductor photo catalytic degradation

	heterogeneous photocatalytic degradation technics based on semiconductors have the
	advantages of low cost and recycle use. In 1972, Fujishima and Honda 83 carried out
	the reaction of photocatalytic cracking water on the n type semiconductor TiO2
	electrode, which made the photocatalytic technology emerged on the historical stage.
	Photocatalysis technology, as a new pollution control method, uses semiconductor
	materials to convert light energy into chemical energy and electric energy. Because of
	its advantages of high efficiency, low cost, complete oxidation and no second
	pollution, it has attracted wide attention from researchers.
	Recently, the commonly researched semiconductors included TiO2 84 , ZnS/Ag 85 ,
	WO3 86 , BiOX 87 , CdS 88 , V2O5 89 , Ag3PO4 90, 91 and so on. Under a certain wavelength
	of light excitation, valence band electrons can be excited into the conduction band,
	and photogenerated electron hole pairs can be generated, following the reaction with
	water, hydroxide, and oxygen, which can generate radicals after (see R7
	Compared with traditional homogeneous AOPs, the formation of radicals need
	the addition of H2O2, Na2S2O8 or iron ions, will easily make the second pollutant. So

-11, Fig 2-4).

As shown in Fig.2-5, similar to

	O application SO4 •-, attention should be paid to the addition amount of persulfate to O O
	S avoid the self-quenching effects. Furthermore, in the activation by iron ions, the molar -O S O O O H O O S O --
	O ration between iron ions and persulfate also is very important because of the high O O
	reaction rate between ferrous ions and sulfate radical (see R2
					(a)	(b)
					Fig.2-5 Structure of persulfate and peroxymonopersulfate
		The generation of SO4 •-	usually derives from the activation of persulfate (PS) by
	heat 70	, UV and transition metals 69, 109, 110	. Among these activation processes,
	Fe 0	/Fe 2+	/Fe 3+	/UV-PS methods are efficient, energy-saving and relatively nontoxicity 69	.
	The main formation of HO 	and SO4 •-	from iron activation route is shown below 69,
	111
	:
	S 2 O 8 2-+Fe 2+ →SO 4 •-+SO 4 2-+Fe 3+	k= 20-27 M -1	s -1	(R2-26)
	S 2 O 8 2-+UV→2SO 4 •-	(R2-27)
	S 2 O 8 2-+heat→2SO 4 •-	•-has been studied long (R2-28)
	time ago, the application of AOPs based on SO4 •-to environmental treatment was 2-n+ (n+1)+ -2-S O +M M SO SO     (R2-29)
	2 8			4	4
	started in recent years. In wastewater treatment, peroxysulphate only adopted as the
	44 source of sulfate radical. Peroxysulphate mainly includes two classes, one is -n+ (n+1)+ 2-SO M M SO     (R2-30)
	2 8 persulfate (PS), and the other one is peroxymonosulfate (PMS). The representative 4 4 2-0 n+ -2-2S O +M M 2SO 2SO     (R2-31)
	When S2O8 2-compound are Na2S2O8, K2S2O8, (NH4)2S2O8 and KHSO5, respectively. H2O2, PS and PMS are both strong oxidizing agent, possessing is activated, it will produce SO4 •-, which will undergo a series of free radical chain reaction (R2-32, 2-33) in aqueous solution, and eventually SO4 •-
	"O-O" perssad in structure. Compared with H2O2 and O3, PS and PMS are more and HO  will coexist in water. The amount of existence of the two radicals is
	stable in low temperature, they can be existed for several weeks in underground determined by the pH value of the solution. Guo-Dong Fang et al. 112 used Electron
	environment. The utilization of PMS was not so much in the previous literatures, so
	Paramagnetic Resonance (EPR) to study the types of radicals existing in the solution
	normally, peroxysulphate usually indicating PS. (NH4)2S2O8 can release ammonia gas of S2O8 2-under different pH conditions. The results show that, in acidic and neutral
	during the application process. The solubility of K2S2O8 is relatively low in water. So conditions (pH = 2-7), is mainly SO4 •-exist; in weak alkaline condition (pH = 8-10),
	with a solubility 549 g L -1 , low-cost Na2S2O8 has obtained the maximum use. SO4 •-and HO  co-existed; in strong alkaline conditions (pH > 11), HO• has become
	the main oxidative species solution. However, because of the weak oxidation of HO 
	under alkaline conditions, the effect of SO4 •-	oxidation system on alkaline wastewater
	treatment is lower than that under acidic or neutral conditions. Same with H2O2, in the

-34, 2-35 and 2-36)

Table 2 -

 2

	Reference	Initial concentration	Water matrix	AOPs	Removal percentage(%)
	Diclofenac
	Perz-Estrada et al.,2005 79	50 mg L -1	Freshwater	Photofenton under solar irradiation	Completely removed in 100 min
				200 mg L -1 TiO2
	Calza et al.,2006 131	15 mg L -1	Distilled water	photocatalytic degradation under	Completely removed in 60 min
				solar irradiation
	Hartmann et al.,2008 132	50 mg L -1	Distilled water	Ultrasonic mg L -1 TiO2 degradation with 100	85% removed in 30 min
	Carbamazepine
	Doll et al., 2005 78	4.2 mg L -1	Lake water with 0.5 mg L -1 NOM	100 mg L -1 TiO2 solar irradiation photocatalytic degradation under	75% removed in 9 min

5 Representative studies dealing with the top 5 pharmaceuticals most commonly treated by AOPs. na: not available

. Bismuth-based catalyst 2-4-1. Reactivity mechanism In

  that be responsible to visible light. Bismuth based oxide has good responsibilities to visible light and relatively high strong oxidation abilities. For example, semiconductor Bi2O3 with a band gap 2.8 eV 147 can adsorb visible light. Furthermore, Bi2O3 is non-toxicity and stable, which deciding that it's a promising catalyst in the future. On the other hand, Zhang and his group 148 synthesized BiOX

order to make catalyst more available to the application of wastewater treatment, many researchers focused on the catalyst ability to absorb solar light. As reported, the ultraviolet light (𝝀 <400 nm) in the solar light is about 3%-5%, and the visible light (400< 𝝀 <800 nm) can account for 47%, and the rest is the infrared radiation 140 . Many catalysts can be only activated by UV and their application is limited for the practical wastewater treatment. Nowadays different researches focused on the new types of photocatalyst, which mainly concentrated on some layered compounds and structural compounds. The representative catalyst is Bi-based materials, such as Bi2O3 141, 142 、BiOX 143 、Bi2O2CO3 144 、Bi2WO6 145 、Bi25VO40 146 . The feature of these semiconductors concludes special layered structure and suitable bad-gap

Table 2 -6. Table 2-6

 2 Synthesis of BiOX catalyst and degradation of pollutant BiOCl/BiOI, Dong 159 et al synthesized BiOCl/BiOI catalyst with different chlorine and iodine molar ratio by precipitation methods at room temperature. The morphology of the catalyst showed flower-like. From the characterization of X-ray powder diffraction (XRD), solid solution was formed between these two catalysts.Under the irradiation of visible light, composite catalyst showed better removal efficiency in degrading NO compared with single catalyst. When the molar ratio of chlorine and iodine was 3:1, it showed the fastest removal rate. Xiao 185 et al also synthesized BiOCl/BiOI composite catalyst by solve-thermal method. Under the Among the composite catalyst, BiOCl plays a major role, and the holes produced on the valence band is the main active species, and the main function of Bi2O3 is the photosensitizer that absorbs visible light.Li et al. 188 synthesized BiOI/Bi2O3 catalyst, under visible light, it has good degradation effect on phenol and four chlorophenol. By studying the two kinds of catalysts, we can find that the good catalytic efficiency is mainly due to the effective separation of electrons and that can effectively separate the photogenerated electron hole pairs and promote the formation of free radicals.Duo et al. 190 also synthesized BiOCl/TiO2 catalyst by hydrolysis at low temperature. Under simulated solar light, the removal efficiency was highly improved towards the degradation of phenol.The main active species were hole and superoxide radicals determined by radicals quenching experiments. This magnetic material can be easily removed from water by magnet, opening a new word, opening a new chapter for the recovery and utilization of the Bi catalyst.Xu et al. 195 synthesized a 3D-structered Bi2WO6-Fe3O4, it showed better degradation abilities than single Bi2WO6 and P25 in removing RhB under the irradiation of visible light. 98.6% of RhB disappeared in 150 min irradiation. By the same way, it also can be recovered effectively from water by external magnet. (7) BiOX/Ag. The deposit of noble metals has always been a big direction in the improvement of the catalyst, and BiOX/Ag/AgX has been widely studied[196][197][198][199][200][201][202][203][204][205] .

	Frference	Synthesis method	Irradiation light	Pollutant
	BOCl
	Yu et al. 165	Low temperature wet chemical	UV	Methyl orange and phenol
	Wang et al. 170	Solve thermal	Visible	Rhodamine B
	Gao et al. 172	Solve thermal	Xenon lamp	Carbamazepine
	Zhang et al., 173	Precipitation	UV	Rhodamine B
	Zhang et al., 161	Hydrolysis	High pressure	Methyl orange
			mercury

and BiOI belongs to p type semiconductor

184

. So the modification and optimization of BiOX mainly concentrated on the composite catalyst, especially p-n types. Now the representative modification is described as follows:

(1) irradiation of visible light, removal rate of Biphenol A (BPA) was much better when treated by composite catalyst, and it was 40 times higher than P25. When the molar ratio of chlorine and iodine was 1:9, it showed the best removal efficiency. This phenomenon mainly can be attributed to the heterojunction which was formed in the composite catalyst that can decrease the combination rate of electron-hole pairs. Wang 186 et al and Kim 171 et al synthesized the similar BiOCl/BiOI composite catalyst, all showing better photocatalytic properties ascribing to its special band structure and great surface area. (2) BiOX/Bi2O3, Chai 187 et al synthesized BiOCl/Bi2O3 composite catalyst, under the visible light irradiation, both BiOCl and Bi2O3 showed very low catalytic efficiency. While, the removal of organic matter has been improved by the composite catalyst, and the organic matter can be completely mineralized and no intermediate products are produced.

holes and reduction of recombination by composite catalysts (3) BiOX/TiO2, Li et al.

189

synthesized BiOCl/TiO2 catalysts by solvothermal treatment, the reason for better catalytic properties mainly can be attributed to the heterojunction formation hierarchical structure and great surface area. Under visible light irradiation, the degradation efficiency of RhBs higher by this 3D flower-like crystal than the 2D plate like BiOCl/ Bi2S3, which can be attributed to its greater surface area and higher separation rate of photo-induced hole and electron pairs. By the same ion-exchange method,

Cao et al. 192

synthesized BiOI/Bi2S3, under visible light activation, it showed excellent degradation of methyl orange, and 4% Bi2S3/BiOI showed the best efficiency among these prepared catalyst due to its same reason with the above description. However, this kind of catalyst did little work to the colorless pollutant like phenol or four chlorophenol. (5) BiOX/Bi2O2CO3,Zhang et al. 144 obtained BiOCl, BiOCl/Bi2O2CO3 and Bi2O2CO3 catalysts by adding different amount to the bismuth nitrate solution. Under simulated solar light, composite catalyst showed better degradation abilities than single catalyst, and it can remove 95% COD of methyl orange solution in 8 hours. visible light was taken as irradiation source, RhB can be removed up to the 90.1% in 40 min. While single BiOI only can remove 50.3% of RhB during the same period.

After recycle for five times, there was no obvious decrease in degradation abilities.

Cheng et al. 203

synthesized AgI/BiOI by precipitation method at low temperature. The fluorescence of composite catalyst was weaker than the single catalyst, indicating its lower recombination rate of photo-induced hole-electron pairs. Under irradiation with visible light, composite catalyst activated better than the single catalyst in removing methyl orange and phenol. Furthermore, the addition amount of AgI played significant roles, and when the addition achieved 20%, it exhibited the best. It was found that AgI stayed stable in the recycle use of AgI/BiOI, showing promising abilities in the practical use of wastewater treatment under visible light.

  . Due to these physicochemical properties of iron in water, the activation of persulfate by Fe 2+ /Fe 0 is more favorable in acidic solution than in neutral and alkaline pH[START_REF] Nie | Degradation of chloramphenicol by persulfate activated by Fe2+ and zerovalent iron[END_REF] . So the development of Fe(II)/Fe(III) organic complexes is essential to improve the removal efficiencies in higher pH solutions.

	Polycarboxylate acids like citrate, oxalate and aminopolycarboxylic acids like
	ethylenediaminetetraacetic acid (EDTA) can form stable water soluble complexes
	with iron in neutral and slightly basic pH solutions, enhancing the dissolution of iron
	in natural water. Moreover, these complexes are photochemically active and efficient
	leading by photoredox process to the production of oxidative species like hydroxyl
	radicals 100, 208, 209 . Among the complexes formed between EDTA and metals,

Fe(III)-EDTA is obviously proved to be the most photosensitive. Different researches revealed that the decay of Fe(III)-EDTA under UV irradiation depends strongly on the wavelength of irradiation and the pH value of environment, as it is reported that the half-life of Fe(III)-EDTA in surface waters ranges from 11.3 min to more than 100 h, which depend on the different light conditions employed. And the intermediate and decay pathways of photodegradation of Fe(III)-EDTA complex depends significantly on the oxygen of the system.

Zhao et al. 210

studied the degradation process of iron oxalate (Fe(III)-oxalate) complex under ultraviolet light and visible light. Unlike most advanced oxidation processes, Fe(III)-oxalate/H2O2/photochemical system can initiate to both reductive and oxidative degradation processes.

Zhou et al. 211

compared the degradation of diethylstilbestrol in water by Fe (III)/UV and Fe (III)-oxalate/UV system. The results

Table 2 -7.

 2

	EDDS is a structural isomer of EDTA, has two chiral centers and exists as three
	stereoisomers, namely [S,S]-EDDS (25%), [R,R]-EDDS (25%) and [R,S/S,R]-EDDS
	(50%), see Fig.2-7. As it is reported in the previous research, [S,S]-EDDS is readily
	biodegradable, while the [R,S/S,R] and [R,R] stereoisomers are less biodegradable 213,
	214 .
	HO 2 C	H	H			H O 2 C	H	H
	HO 2 C	N H	HO 2 C N	H	CO 2 H	H O 2 C	N H	N H C O 2 H C O 2 H

Table 2 -7 Chemical and physical properties of EDDS ethylenediamine-N,N'-disuccinic acid

 2

	CAS number	20846-91-7
	Molecular formula	C10H16N2O8
	Molecular weight	292.24 g/mol
	Melting point	516.68 °C at 760 mmHg
	Boiling point	220-222 º C
	λmax	none
	Intensity	1.44 g/cm³ (25℃)
	logKow	-5.44 (KowWin est.)
	Solubility in water	slightly soluble
	Flash point	38 °C

  •-, 10 mM Na2S2O8 and different concentrated p-HPA solutions were set under pulse irradiation. Sulfate radical decay was followed at 450 nm corresponding to the maximum absorption of this species.

11 eV 164.42 eV 274 276 278 280 282 284 286 288 290 292 294 7.6x10 4 7.8x10 4 8.0x10 4 8.2x10 4 8.4x10 4 8.6x10 4 8.8x10 4 Count (a.u.) Banding Energy (eV) C 1s 284.6 eV 281.39 eV

		1.05x10 5								1.1x10 5
				159.							Bi 4f
											O 1s
		9.00x10 4								1.1x10 5	531.85 eV
		7.50x10 4								1.1x10 5
	Count (a.u.)	4.50x10 4 6.00x10 4								Count (a.u.)	9.5x10 4 1.0x10 5	530.61 eV	532.83 eV
											9.0x10 4
		3.00x10 4
											8.5x10 4
		1.50x10 4
											8.0x10 4
			152	154	156	158	160	162	164	166	168	170
											526	528	530	532	534	536	538
						Banding Energy (eV)
			9.4x10 4								Banding Energy (eV)
											Cl 2p
									197.95 eV
			9.2x10 4
			9.0x10 4
			8.8x10 4
		Count (a.u.)	8.2x10 4 8.4x10 4 8.6x10 4							199.60 eV
			8.0x10 4
			7.8x10 4
						Banding Energy (eV)

Table 4 -

 4

		1 physicochemical parameters of BiOCl0.75I0.25 catalyst
		BET surface	Average pore	Total pore density		Band-gap
	Parameters	area (m 2 g -1 )	diameter (nm)	(cm 3 g -1 )	Morphology	(eV)
		34.93	16.99	0.1483	Flower-like	2.12

  •-(R4-5) 239 . For the third reason, G.

Fang et al. has demonstrated the reaction of O2

Table 4 - 4

 44 The intermediate products of p-HPA and their proposed structures

	Products	Tret (min) a	Main fragment ions	Structure
	1	6.415	124; 95, 77
	2	6.540	122; 93, 65

a Retention time.

Table 5 - 1

 51 Second order constant kp-HPA,HO• and kp-HPA,SO4•-at two different pH

		pH 2.5	pH 8.5
	kp-HPA•HO• (M -1 s -1 )	(2.2±0.12)×10 10	(7.27±0.30)×10 9

  • /O2 •-. This hypothesis was deduced from the literature(Miralles-Cuevas et al. 2014) [START_REF] Miralles-Cuevas | Pharmaceuticals removal from natural water by nanofiltration combined with advanced tertiary treatments (solar photo-Fenton, photo-Fenton-like Fe(III)-EDDS complex and ozonation)[END_REF] , EDDS• can react with anions in aqueous solution like hydroxyl to form related radicals. Furthermore, the existence of S2O8 2-can consume electrons and O2 •-to form sulfate radicals (R5-13-R5-14)240, 244 . While it's well known that HO2 • /O2 •-radicals are particularly important because they can accelerate the Fe(III)/Fe(II) cycle and so the reactivity as it is described in the paper ofHuang et al.. 111 Moreover, as it is demonstrated in the paper ofWu et al. (2014) [START_REF] Ng | A novel application of anaerobic bio-entrapped membrane reactor for the treatment of chemical synthesis-based pharmaceutical wastewater[END_REF] the degradation efficiency is mainly affected by the photochemical activation abilities of different formed of Fe(III)-EDDS complex and the soluble Fe(III) concentrations

	in solution 217 . With pH increase, Fe(III)-EDDS is less efficient photochemicaly, and
	the precipitation of Fe(III) in aqueous solution is stronger. Comparison between Fe 3+
	and Fe(III)-EDDS also was conducted, as shown in Fig.5-10B. At pH 2.0,
	Fe 3+ /Na2S2O8 (500 µM)/UV system shows a fast and gradual degradation of p-HPA,
	and p-HPA can be removed completely in 45 min.
	EDDS•+S 2 O 8 2-→S 2 O 8 •-+Product	(R5-12)
	e aq -+S 2 O 8 2-→SO 4 2-+ SO4 •	(R5-13)
	S 2 O 8 2-+O 2 •-→SO 4 •-+ SO 4 2-+O 2	(R5-14)

  10 M -1 s -1 .

	HO  +ACTP→product	(R6-17)
	SCN • +SCN -→SCN 2 •-	(R6-18)
	H 2 O 2 +hν→2HO 	(R6-15)
	HO  +SCN -→HO -+SCN •	(R6-16)

  9 M -1 s -1 when competing with BPA. Homogeneous AOPs than involved Fe(III)-EDDS and heterogeneous AOPs that used BiOCl0.75I0.25 both had good response to simulated solar light, and can remove ACTP effectively. Fe(III)-EDDS induced degradation was fast, efficient and effective, while it has the risk to expose some ions like Fe(III) or SO4 2-to water. BiOCl0.75I0.25 induced heterogeneous degradation was relatively mild and clean, however, recycling problem need to be solved before its practical application. 158. Hu, J.; Fan, W.; Ye, W.; Huang, C.; Qiu, X., Insights into the photosensitivity activity of BiOCl under visible light irradiation. Appl. Catal. B 2014, 158-159, 182-189. 159. Dong, F.; Sun, Y.; Fu, M.; Wu, Z.; Lee, S. C., Room temperature synthesis and highly enhanced visible light photocatalytic activity of porous BiOI/BiOCl composites nanoplates microflowers. J Hazard Mater 2012, 219-220, 26-34. 160. Wang, P.; Wu, Y.; Shi, J.; Liu, D.; Dong, W., Preparation of carbon-supported Bi/Ti composites and its catalytic activity under solar irradiation. Appl. Surf. Sci. 2014, 292, 1077-1082. 161. Zhang, K.; Liu, C.; Huang, F.; Zheng, C.; Wang, W., Study of the electronic structure and photocatalytic activity of the BiOCl photocatalyst. Appl. Catal. B 2006, 68, (3-4), 125-129. 162. Wang, X.; Chen, H.; Li, H.; Mailhot, G.; Dong, W., Preparation and formation mechanism of BiOCl0.75I0.25 nanospheres by precipitation method in alcohol-water mixed solvents. Journal of Colloid and Interface Science 2016. Appl. Surf. Sci. 2013, 284, 497-502. 166. Huo, Y.; Zhang, J.; Miao, M.; Jin, Y., Solvothermal synthesis of flower-like BiOBr microspheres with highly visible-light photocatalytic performances. Appl.

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163. Liu, Q. C.; Ma, D. K.; Hu, Y. Y.; Zeng, Y. W.; Huang, S. M., Various bismuth 164. Shi, X.; Chen, X.; Chen, X.; Zhou, S.; Lou, S., Solvothermal synthesis of BiOI hierarchical spheres with homogeneous sizes and their high photocatalytic performance. Materials Letters 2012, 68, 296-299. 165. Yu, J.; Wei, B.; Zhu, L.; Gao, H.; Sun, W.; Xu, L., Flowerlike C-doped BiOCl nanostructures: Facile wet chemical fabrication and enhanced UV photocatalytic properties. 168. Zhang, K.-L.; Liu, C.-M.; Huang, F.-Q.; Zheng, C.; Wang, W.-D., Study of the electronic structure and photocatalytic activity of the BiOCl photocatalyst. Appl.

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Acknowledgement

This project has been carried out within the cooperation program between Fudan University and Clermont Auvergne University. One part of my work was performed in the Department of Environmental Science and Engineering of Fudan University directed by Prof. Wenbo Dong and Dr. Hongjing Li, the other part was carried out in the Institut de Chimie de Clermont-Ferrand (ICCF) UMR 6296 led by Dr. Gilles MAILHOT and Marcello BRIGANTE. I would sincerely like to thank them for providing the good experimental conditions, supervision, and help. I would like to express my thanks to Prof. Wenbo Dong and Dr. Hongjing Li in Fudan University during my whole study period. When I entered Fudan University in 2013 at the first time, I knew little about my profession knowledge. They showed me how to start scientific research, including searching literatures and doing experiments. They help me design my project plan, and discussed with me frequently. Their professional spirit gave me a lot of encouragement. Without them, I could not get this chance to study abroad in Clermont-Ferrand. They also paid special attention on my life and family. I want to say thank you to Dr. Gilles MAILHOT and Dr. Marcello BRIGANTE who directed my experiments in France and help me to finish my papers and thesis. They gave me a lot of help in my study and life during my stay in France. Both of them always gave me helpful suggestion and idea for the whole study, their hard-working attitude and rich experience gave me deep impression and influenced me in my whole life. I also want to thank sincerely Mr. Zhijian ZHANG, Miss Yanlin WU, Mr. Minghua NIE, Mr. Peng WANG, Miss Ping GONG, Miss Jin SHI, Miss Yicen XUE, Miss Pingping ZHAI, Miss Hongche CHEN, Miss Yahong SHI, Mr. Yankun LIU, Mr. Wenlong BI, Mr. Qi LIU, Mr. Shiyu LI, and Miss Juan ZAHO, who are my colleagues in the laboratory and friends in Fudan Univesity. Particularly, I have to thank sincerely to Angelica, an Italian girl, my best friend in France. Furthermore, I want to thank Wenyu HUANG, Lei ZHOU, Jinlong ZHA, Shengnan LI, Miao WANG, Jinpeng WANG, they gave me such a happy time during my stay in France. Thanks for the fund support of China Scholarship Council (CSC) affiliated with the Ministry of Education of the P. R. China. Thanks also go to the Education Service of China Embassy in Paris. Finally, I would thank my parents, they gave me a lot of care, support and encourage. I can't finish my Ph.D study without them. I thank my husband and my baby, they are my spiritual pillar during my whole life. I love you forever.