Président: Hui Zhang

Feng Wu

Gilles Mailhot

OXIDATION OF ORGANIC COMPOUNDS BY OXYSULFUR RADICALS IN THE PRESENCE OF TRANSITION METAL IONS AND SULFITE

Oxidative of organic compounds by oxysulfur radicals in the presence of transition metal ions and sulfite

I 1

Abstract

In recent years, more and more refractory and toxic organic compounds are detected in wastewater. Many of these organic pollutants are hardly degraded by conventional water treatments. Sulfate radical based advanced oxidation process (SR-AOPs), have emerged as a promising method in the field of oxidative decontamination of polluted water. Past studies focused on this SR-AOPs with peroxydisulfate (PS) or peroxymonosulfate (PMS) as oxidants, especially the systems involving transition metals and oxidants (i.e. Fe(II)/PS system, Ni(II)/PMS system and Co(II)/PMS system), which has been confirmed that SO4 -could have some advantages over HO  in the decontamination of wastewater containing organic pollutants. In this dissertation, oxysulfur radicals including sulfite radical SO3 -, sulfate radical SO4 -, peroxymonosulfate radical SO5 -produced by transition metal ions such as Cr(VI), Co(II), Fe(III) activated sulfite are used to degrade organic compounds. The removal efficiency, the oxidation mechanism are examined, and the role of sulfur species are elucidated. The main content and results consist of:

(1) Cr(VI)-S(IV)-O2 system containing sulfite ions that rapidly and simultaneously reduces Cr(VI) and oxidize organic pollutants in the presence of oxygen in aqueous solutions is investigated. This Cr(VI)-S(IV)-O2 system contains the initiator Cr(VI), the reductant S(IV), and the oxidant O2, which produce oxysulfur radicals (mainly SO4 -and SO5 -) and hydroxyl radicals (HO • ). The Cr(VI)/S(IV) molar ratio, pH, and oxygen content play important roles in the entire reaction system. Acidic conditions (pH 3.0) facilitates degradation of organic compounds and reduction of Cr(VI) as well. In addition, experiments of rapid degradation of several kinds of organic pollutants are also conducted.

Preliminary results show that the removal rates of the analogs of phenols or aromatic amines in this Cr(VI)-S(IV)-O2 system have a relationship with the electronic parameters (Hammett constant, σ) of the substituted groups. Thus, the Cr(VI)-S(IV)-O2 system, provides an excellent strategy of "waste control by waste" for removing multiple industrial contaminants.

(2) The efficiency Cobalt (II) (Co(II)) for the activation of sulfite ions following the oxidation of paracetamol used as model contaminants is investigated. Physicochemical parameters that can impact the paracetamol degradation (pH, initial paracetamol concentration, Co(II)/S(IV) molar ratio, oxygen concentration) and contribution of various radicals are investigated in order to elucidate the chemical mechanism. Main results show that the pH is a key factor controlling the efficiency in the system Co(II)/Sulfite. Higher efficiency is observed for pH between 9.0 and 10.0. Increasing S(IV) concentrations, until 1 mM, slightly promoted the degradation of paracetamol. In fact, an excess of sulfite ions inhibits the reaction through the scavenging of SO4  -and SO5  -. Moreover, degradation efficiency drastically decreases from ~ 85% to less than 5% in absence of oxygen. SO4  -is confirmed to be the main oxidant responsible for the paracetamol degradation. For the first time the second order rate constant between SO4  -and paracetamol (1.33 ± 0.79 × 10 9 M -1 s -1 (at pH 5) and 6.14 ± 0.99 × 10 8 M -1 s -1 (at pH 11.0)) is determined. Moreover, radical-scavenging experiments also suggest the possible implication of SO5  -.

(3) The efficiency of Fe(III)-S(IV) system has been investigated using aniline as pollutant model compound in water. Chemical kinetics, influencing factors, and mechanism of aniline oxidation are examined with emphasizing the contribution of different oxysulfur radicals (mainly SO4  -and SO5  -). Our results show a significant enhancement of aniline oxidation at pH 4.0 with 1.0 mM of S(IV) and 0.1 mM Fe(III). Moreover, degradation efficiency drastically decreases to 10% in absence of oxygen indicating the relevance of oxygen in this process. Through competition kinetic and radical scavengers experiments we show that SO5  -is responsible for about 60% of the aniline oxidation in the Fe(III)-S(IV) system under our work typical conditions. For the first time we determine the second order rate constants between SO5 -and aniline (5.8 ± 0.6 × 10 6 M -1 s -1 (at pH 3) and SO4 -and aniline 7.7 ± 0.5 × 10 9 M -1 s -1 (at pH 3.0)). Moreover, sequential experiments with addition of sulfite improve the oxidation efficiency.

(4) The photodegradation of paracetamol (PARA) has been investigated in Fe(III)-S(IV)-O2 system in the absence and presence of UV irradiation. The results indicate that pH, Fe(III)/S(IV) concentration and oxygen content have an important impact on the disappearance of PARA. To achieve the highest efficiency of PARA degradation, the higher S(IV) concentration should be chosen. Irradiation enhances the PARA degradation through the pH decrease. In fact, the optimal pH range for the degradation of PARA in Fe(III)-S(IV)-O2 system is found to be around pH 4.0. Combining the radical scavenger and the ESR experiments, we show that the primary degradation pathway of PARA is the reaction with SO4 -. Experiments are also conducted to examine whether other transition metal ions such as Cr(VI), Co(II) and Cu(II) can react with S(IV) at initial pH of 7.0 under irradiation, and results show that PARA can only be degraded in Cr(VI)-S(IV)-O2 system with UV irradiation, Co(II) and Cu(II) show no efficiency on PARA degradation under such environment. In conclusion, this work indicate that UV irradiation could be a plausible pathway for PARA degradation by Fe(III)-S(IV)-O2 system with initial pH around 7.0. This reactivity makes the Fe(III)-S(IV)-O2-hv system a more powerful system to deal with the wastewater at near-neutral pH values. The increase of chemical product use and industries in the recent years has strongly enhanced the presence of refractory, toxic, carcinogenic and mutagenic organic compounds in the aquatic compartment. Organic compounds include dyes, detergents, pharmaceuticals, personal care products, pesticides, carboxylic acids and aromatic compounds [1][2][3][4][5][6][7]. The presence of those organic compounds in water poses serious threat to water availability in some regions all over the world [8][9].The frequent occurrence of pollutants in the aquatic environment as well as in drinking water has raised a concern about their potential impact on environmental and public health. Human exposure to these organic pollutants causes damages to nervous central system, blood system and primary organs like livers, lungs, kidneys etc even at very low concentration [10][11]. Moreover, many organic compounds cannot be eliminated by conventional physical separation methods and biological processes due to their strong stability and toxicity [12]. Therefore, more powerful wastewater treatment methods are required.

Advanced oxidation processes (AOPs) such as Fenton, photo-Fenton, ozone oxidation, sonolysis and photo-catalytic oxidation have been proved to be effective for degradation of many organic/inorganic compounds from aqueous medium [13][14][15][16][17][18]. Among the AOPs, those which produce hydroxyl radicals (OH  ) are the most successful, and attracted attention in the field of oxidative decontamination of polluted waters and soils. In fact, HO  has a strong oxidation potential (E 0 = 2.8V), is highly reactive and has a non-selective reactivity [19][20].

Fenton process is based on an electron transfer between hydrogen peroxide (H2O2) and a homogeneous metal catalyst (Fe 2+ ) [21]. In Fenton process, hydrogen peroxide is catalyzed by ferrous ion to produce HO  [22].

Fe 2+ + H2O2 → Fe 3+ + HO -+ HO • 1-1

The use of Fenton reaction to archive the degradation of several refractory organic compounds such as chlorophenols, chlorobenzene, nitrophenols, has been demonstrated to be an efficient method [23][24][25][26]. Furthermore, Fenton process can be improved by using UV radiation and hydrogen peroxide (H2O2), which is so-called photo-Fenton process as shown in eqs 1-2 and 1-3 [27][28][29];

H2O2 + h → 2 HO  1-2

Fe 3+ + H2O + h → HO  + Fe 2+ + H + 1-3

Muruganandham at al. [27] investigated the degradation efficiency of Reactive Orange 4 using Fenton and photo-Fenton systems. In their work the disappearance of dye reaches 79.89 % and 95.5% after 40min using Fenton and photo-Fenton processes respectively. In addition, besides photo-Fenton, a combination of electrochemical and Fenton process, namely electro-Fenton process also has been successful applied in wastewater treatment and achieving a good degradation and mineralization result [30][31][32].

However, for Fenton system, the disadvantage is related to the high Fe 2+ concentration and the formation of Fe sludge which limit its application [33].

Nowadays, an efficient method for chemical oxidation based on the sulfate radical (SO4 -) has gained widespread attention (so called Advanced oxidation processes based on sulfate radical (SR-AOPs)) [34][35]. SO4 -is a strong oxidant with relatively high standard redox potential (E 0 = 2.6 V vs NHE) that is closed to the value for hydroxyl radical [36]. In the recent decades, studies have focused on the application of SO4 -for the removal of organic contaminants.

SO4 -can react via electron transfer, by addition to C-C double bonds and H-abstraction [37,38], thus, it is able to oxidize a large number of pollutants such as dyes, pesticides, antibiotics and aniline in water [39][40][41][42][43][44]. SO4 -can be produced from persulfate (PS) or peroxymonosulfate salts (PMS) by a variety of approaches such as photolysis and thermal activation [45][46][47][48]. The addition of transition metals is recognized to be a viable way to achieve PS/PMS activation [49]. Co(II) and Fe(II)/Fe(III) are the most commonly used due to their natural presence in environmental media [50][51].

In the atmospheric liquid phase, Fe(II/III) catalyzes the conversion of S(IV) to S(VI) in the presence of oxygen. Many investigations have extrapolated detailed mechanisms under various conditions, since it is very important for the generation of acid rain [START_REF] Berglund | Manganese-catalysed autoxidation of dissolved sulfur dioxide in the atmospheric aqueous phase[END_REF][START_REF] Hoffmann | Kinetics and Mechanisms of the Catalytic Oxidation of Dissolved Sulfur Dioxide in Aqueous Solution: An Application to Nighttime Fog Water Chemistry[END_REF][START_REF] Brandt | Kinetics and mechanism of the iron(III)-catalyzed autoxidation of sulfur(IV) oxides in aqueous solution. The influence of pH, medium and aging [J][END_REF][START_REF] Kraft | ChemInform Abstract: Kinetics and Mechanism of the Iron(III)-Catalyzed Autoxidation of Sulfur(IV) Oxides in Aqueous Solution. Part 1. Formation of Transient Iron(III)-Sulfur(IV) Complexes [J][END_REF][START_REF] Zuo | Effects of Monochromatic UV-Visible Light and Sunlight on Fe(III)-Catalyzed Oxidation of Dissolved Sulfur Dioxide [J][END_REF][START_REF] Sun | Activation of Manganese Oxidants with Bisulfite for Enhanced Oxidation of Organic Contaminants: The Involvement of Mn(III) [J][END_REF]. On the basis of the mechanism of catalysis by iron, much attention has also been paid to flue gas desulfurization. Expect for iron, recently, sulfite ions were also found to react with transition metals such as Ni(II) and Mn(III) to generate SO4 -for water decontamination [START_REF] Sun | Activation of Manganese Oxidants with Bisulfite for Enhanced Oxidation of Organic Contaminants: The Involvement of Mn(III) [J][END_REF][START_REF] Carvalho | Further studies on the synergistic effect of Ni(II) and Co(II) ions on the sulfite induced autoxidation of Cu(II) penta and hexaglycine complexes [J][END_REF].

Furthermore, in order to enhance the efficiency of Fe(II)-S(IV) system, UV-vis irradiation was employed [START_REF] Zhang | Enhanced Decolorization of Orange II Solutions by the Fe(II)-Sulfite System under Xenon Lamp Irradiation[END_REF].

In my PhD work, transition metal ions such as Cr(VI), Co(II), Fe(III) were chosen to react with sulfite to promote the organic compounds degradation in aqueous solution. In order to apply the transition metal ions-S(IV) systems at near-neuter pH, UV-vis irradiation was used to test the feasibility. In all those systems, the activation mechanism and contribution of reactive oxygen and sulfur species were elucidated.

research goals and objective

The objective of this work is to study the degradation of model pollutants from water using transition metal ions-S(IV) systems, including Cr(VI)-S(IV)-O2 system, Co(II)-S(IV)-O2 system, Fe(III)-S(IV)-O2 system and Fe(III)-S(IV)-O2-hv system. The optimum degradation conditions are identified and the activation mechanisms are examined and the role of sulfur species are elucidated. To meet these goals, the following specific objectives for every individual system are defined:

1, Experiments are conducted to investigate the parameters on the rapidly and simultaneously reducing Cr(VI) and oxidizing organic pollutants in the presence of oxygen in aqueous solutions in Cr(VI)-S(IV)-O2 system. The species of oxidants and the contributions of reactive radical species are monitored, then the mechanism of the system can be preliminarily probed.

2, Experiments are conducted to investigate the parameters on the paracetamol degradation in alkaline solution in Co(II)-S(IV)-O2 system. The activation mechanism and contribution of reactive oxygen and sulfur species are elucidated.

3, In Fe(III)-S(IV)-O2 system, experiments are conducted to investigate the parameters on Aniline oxidation. The second-order rate constants between aniline and SO4 -/SO5 -are measured, and the contribution made by SO4 -and SO5 -are calculated. The role of oxygen is emphasized.

4, Experiments are conducted to verify the paracetamol degradation in Fe(III)-S(IV)-O2-hv system at near-neutral pH by UV irradiation. The parameters and oxidant responsible for the organic pollutants degradation are examined.

Advanced oxidation process (AOPs)

Advanced oxidation processes (AOPs) appear as one of the most promising methods for the treatment of soils and waters contaminated by organic compounds [1]. AOPs employ high oxidation-potential sources to produce the primary oxidant species, hydroxyl radical (HO  ), an unselective and powerful oxidizing species. HO  reacts with organic compounds with a second order rate constant in the range of 10 8 -10 10 M -1 s -1 , yielding to CO2, H2O and, eventually, inorganic ions as final products [2][3][4][5]. The advantage of AOPs over all chemical and biological processes is that the AOPs is totally ''environmental-friendly" as they neither transfer pollutants from one phase to the other (as in chemical precipitation and adsorption)

nor produces massive amounts of hazardous sludge [6][7][8]. Due to these characteristics, numerous works have been done to investigate the applications of AOPs to treat different types of contaminated waters. Among the various AOPs proposed in the literature, ultraviolet photolysis (UV), UV/H2O2, Fenton, photo-Fenton process, photocatalysis are the most common and efficient oxidation technologies which have been applied successfully in wastewater treatment. The mechanism of these methodologies along with equations is presented as follows:

(1) Photolysis/ ozonolysis O3 is widely used in water treatment as disinfectant and oxidant. The decomposition of aqueous ozone in water is initiated by its reaction with hydroxide ion (eqs 2-1), and this reaction leads to the production of free radicals that propagate the decomposition process by chain of radical reactions and produce hydroxyl radical. The purpose of introducing UV radiation in the ozonation process is to enhance the ozone decomposition, yielding more free radicals for reaching a higher oxidation rate 2-2) [9]: H2O2 + hv→ 2HO  2-3 When compared the system H2O2 and H2O2/UV, it was found that UV increased the percentage of degradation to a large extent [14][15].

2O3 +H2O2 → 2 HO  + 3O2 2-1 O3 + hv + H2O →HO  + other species 2-2 (
(

3) Fenton process

The generally accepted mechanism of the Fenton process proposes that HO  are produced in accordance with eqs 2-4 [16][17][18][19][20]; while the catalyst is regenerated through eqs 2-5, or from the reaction of Fe 3+ with intermediate organic radicals (eqs (2-6)-(2-7)) [21][22], as ferrous ions are consumed more rapidly than they are regenerated, hence high concentration of Fe 2+ is required in Fenton process [23].

Fe 2+ + H2O2 →Fe 3+ + HO -+ HO  2-4 Fe 3+ + H2O2 →Fe 2+ + HO2  + H + 2-5 RH + HO  → H2O + R  2-6 Fe 3+ + R  →R + + Fe 2+ 2-7 (

4) Photo-Fenton

Photo-Fenton means when using UV-vis light irradiation, the photolysis of Fe 3+ aqua-complexes (Fe(OH) 2+ ) allows to improve the Fe 2+ generation and the occurrence of Fenton reactions in the presence of H2O2 (eqs.2-8) [24].

Fe(OH) 2+ + hv → HO • +Fe 2+ 2-8 (

5) Photocatalysis

Photocatalysis, which makes use of the semiconductor metal oxide such as TiO2, ZnO, CdS and NiO as catalyst is an extensively studied field of AOPs [25][26][27]. Among them, TiO2 in the anatase form is proved to be the most appropriate one due to its characteristics such as high photoactivity, chemical inertness, non-toxic, low cost and easy to obtain [28]. The initiating procedure of the photocatalytic reaction is the absorption of the radiation with the formation of holes (h + ) in valence band and electrons (e -) in conduction band in femtosecond timescale (eqs 2-9)

TiO2 + hv →e -+ h + 2-9
In water, oxidation reactions can take place to form reactive HO  (eqs. 2-10 and 2-11) [29].

During the photocatalytic process, some other reactive radicals like superoxide radical anion (O2 -) are also formed (eqs. 2-12) [30]. According to Eqs. (2-13)-(2-15)), O2 -can lead to the production of HO • [31].

TiO2 (h + ) + H2O →TiO2 + H + + HO  2-10 TiO2 (h + ) + HO -→TiO2 + HO  2-11 O2 + e -→ O2 - 2-12 O2 -+ H + →HO2 - 2-13
TiO2(e -)+ H + + HO2 -→TiO2 + H2O2 2-14

TiO2(e -)+ H2O2 →TiO2 + HO  + HO - 2-15

Different oxidation technologies of AOPs are powerful in wastewater treatment of organic pollutants. However, there are still some disadvantages which limit its application, the worst one belongs to the narrow pH range limitation (pH 2.5-3.5) more particularly for AOPs using iron species. In fact, iron ions are precipitated at higher pH values. Another AOPs called advanced oxidation processes based on sulfate radical (SR-AOPs) is a Fenton-like system, which can generate sulfate radical (SO4 -) to achieve the organic compounds degradation.

SO4 -can be effectively produced from persulfate (PS) or peroxymonosulfate salts (PMS) by a variety of approaches such as photolysis, radiation, thermal activation and so on.

Advanced oxidation processes based on sulfate radical (SR-AOPs)

PMS-based advanced oxidation processes

Potassium peroxymonosulfate (KHSO5, KHSO4. K2SO4, PMS, the structure shown in Figure 2-1) commercially available as oxone (HSO5 -, PMS), is a widely used as oxidizer in a variety of industrial and consumer applications; specifically, it is sold as a non-chlorine shock-oxidizer for swimming pools and spas. Peroxymonosulfate alone, similarly to hydrogen peroxide [32][33], is not as an efficacious disinfectant as active chlorine compounds. If activated using a catalyzer or photochemically, the generated radicals are very energetic and effective oxidizing agents and it may be used as disinfecting agents. PMS in water solutions is relatively stable in a way that only 5% of active oxygen dropped in 3 days [37]. Moreover, PMS is stable at pH less than 6 and at pH = 12. Minimum stability of PMS is observed at pH of 9 in which concentrations of HSO5 -and SO5 2-are equal (pKa= 9.4) [37][38][39].

Figure. 2-1. The molecular structures of oxone salt

Although PMS is thermodynamically a strong oxidant, its direct reaction with the majority of the pollutants is too slow so that activation is required. From many previous studies, PMS was found to have great potential in water treatment activated by transition metals, especially cobalt, or can be coupled with UV radiation to form SO4 - [40][41].

(

1) PMS activated by transition metal ions in homogeneous systems

Transition metals have been extremely used to produce free radicals for the degradation of organic compounds. Early in 1956, Ball and Edwards reported the decomposition of peroxymonosulfate by transition metal ions, cobalt ions, for the first time [42]. Decades after, Anipsitakis, Dionysiou and other researchers employed cobalt ion to activate PMS to generate sulfate radical for the organic contaminants degradation [34,[43][44][45][46], the general reaction is shown as follows (eqs 2-16) [34]:

Co 2+ + HSO5 -→ Co 3+ + HO -+ SO4 - 2-16
The study reported by Anipsitakis and Dionysiou compares the cobalt/peroxymonosulfate (Co/PMS) process with the traditional Fenton process in the dark, over the pH range 2.0 to 9.0 with and without phosphate and carbonate buffers. 2,4-dichlorophenol, atrazine, and naphthalene are used as model contaminants. The results show that Co/PMS process is more efficient than the Fenton process for the degradation of 2,4-dichlorophenol and atrazine.

Furthermore, when the experiments are conducted at high pH values, the efficiency of the Fenton process decreases, but the reactivity of the Co/PMS system is sustained at high pH values. For the degradation of naphthalene, Fenton process demonstrates higher degradation efficiencies at acidic pH, and Co/PMS system is showed more efficient at neutral or even higher pH values. For the effect of mineralization, as total organic carbon removed, also Co/PMS system is proven to be more efficient than the Fenton system [34].

Expect for Co(II), several other kinds of transition metal ions such as Ag(I), Ce(III), Fe(II), Fe(III), Mn(II), Ni(II), Ru(III), and V(III) have been also tested for the activation of PMS. The generated oxidants in those system were also identified as sulfate (SO4 -), peroxymonosulfate (SO5 -), and hydroxyl radicals (HO  ) [35].

M n+ + HSO5 -→ M (n+1)+ + HO -+ SO4 - 2-16
For V(III)/PMS system, sulfate radicals as well as hydroxyl radicals (at a lesser extent) are generated according to eqs 2-17 and 2-18:

V 3+ + HSO5 -→ V 4+ + HO -+ SO4 - 2-17 V 3+ + HSO5 -→ V 4+ + HO • + SO4 - 2-18
As for vanadium, the situation is much more complicated than other metal ions. The interaction of Ce(III) with PMS leads to the formation of a metal-sulfate radicals complex, and the rate-determining step of the Ce(III)/KHSO5 interaction is suggested to be eqs 2-19:

Ce 3+ + HSO5 -→ [Ce IV (SO4  )] 3+ + HO - 2-19
Caged sulfate radicals also can be generated when using Mn(II) and Ni(II) to activate PMS according to eqs 2-20 and 2-21, and this explains in part the low reactivity of these two couples:

Mn 2+ + HSO5 -→ [Mn III (SO4  )] 2+ + HO - 2-20 Ni 2+ + HSO5 -→ [Ni III (SO4  )] 2+ + HO - 2-21
In Fe(III)/PMS system, Fe(III) is an electron acceptor (oxidant) would decompose peroxymonosulfate radical as shown in eqs 2-22:

Fe 3+ + HSO5 -→ Fe 2+ + H + + SO5 - 2-22
These results indicate that PMS can be activated by different metal ions, and generating oxidants which have great potential in water treatment.

(

2) PMS activated by metal oxides in heterogeneous systems

Transition metal ions are usually applied in wastewater treatment in homogeneous systems.

However, there are some disadvantages that limit the application of homogeneous catalyst [47][48][49][50]. For heterogeneous catalyst, most of which are mesoporous catalysts and play their role by the reactions occur at the surface, it can be considered for decomposition of PMS.

Among heterogeneous catalysts, cobalt oxides such as CoO, Co2O3 and Co3O4 are usually used to activate PMS to degrade various pollutants [47]. Co3O4 nano-particle exhibit excellent performance in PMS decomposition due to presence of Co(II) and Co(III) in molecule structure [51][START_REF] Berglund | Manganese-catalysed autoxidation of dissolved sulfur dioxide in the atmospheric aqueous phase[END_REF][START_REF] Hoffmann | Kinetics and Mechanisms of the Catalytic Oxidation of Dissolved Sulfur Dioxide in Aqueous Solution: An Application to Nighttime Fog Water Chemistry[END_REF]. Various metal oxides (MgO, ZnO, Al2O3, ZrO2, TiO2) have been considered

as support for cobalt species in order to activate PMS. Take MgO for example, the presence of hydroxyl groups on the surface of MgO formed CoOH + intermediate which could promote decomposition of PMS. The order of activity of supports was is MgO > ZnO > TiO2 > ZrO2 > Al2O3 .

(

3) PMS activation by UV irradiation

The use of UV light can lead to the generation of SO4 -and HO  radicals through the photolysis of the oxidants according to the eqs 2-23 [START_REF] Carvalho | Further studies on the synergistic effect of Ni(II) and Co(II) ions on the sulfite induced autoxidation of Cu(II) penta and hexaglycine complexes [J][END_REF][START_REF] Zhang | Enhanced Decolorization of Orange II Solutions by the Fe(II)-Sulfite System under Xenon Lamp Irradiation[END_REF];

HSO5 -h𝑣 → HO • + SO4 - 2-23
The advantage of activation PMS by UV is due to its benign and economic properties for the purpose of peroxide activation via breaking of the chemical bonds [60-61]. The investigation of the parameters which affect the quantum yields of SO4 -and HO  formation, shows that UV wavelength has significant influence on the generation of those radicals. The quantum yields of SO4 -decreased with increase of UV wavelength in the range of 248 to 351 nm [62],

and the maximal quantum yield can be gained at the wavelength of 254 nm. For this reasons, 254 nm is usually used for PMS activation [START_REF] Beltran | Ozone-UV radiation-hydrogen peroxide oxidation technologies[END_REF][START_REF] Getoff | Purification of drinking water by irradiation. A review [J[END_REF].

As an efficient AOPs method, the application of PMS/UV treatment is very beneficial not only for the complete elimination of the organic compounds, but also for the achievement of full oxidation (mineralization) [START_REF] Carvalho | Further studies on the synergistic effect of Ni(II) and Co(II) ions on the sulfite induced autoxidation of Cu(II) penta and hexaglycine complexes [J][END_REF][START_REF] Sharma | Oxidative removal of Bisphenol A by UV-C/peroxymonosulfate (PMS): Kinetics, influence of co-existing chemicals and degradation pathway[END_REF][START_REF] Verma | Application of UV-C LED activated PMS for the degradation of anatoxin-a[END_REF].

(4) Thermal activation of PMS Temperature increase is an effective way to activate the PMS leading to the formation of SO4 -and HO  [67-68] (eqs 2-24). In the heat/PMS system, temperature is the critical factor that affect the quantum yields of SO4 -and HO  formation.

HSO5 -△ → HO • + SO4 - 2-24
Furthermore, HO  could also be formed via the reaction of SO4 -and H2O, during the heat process based on the eqs. 2-25

SO4 -+ H2O △ → HO • + SO4 2-+ H + 2-25
Even though the reaction rate of this reaction is too slow to be important in most reaction systems [START_REF] Yang | Degradation efficiencies of azo dye Acid Orange 7 by the interaction of heat, UV and anions with common oxidants: persulfate, peroxymonosulfate and hydrogen peroxide [J][END_REF], but higher temperature could significantly accelerate its reaction rate. In addition, at certain temperature, pH has also an influence on the transformation of SO4 -to HO  . SO4 -was the predominant radical at pH < 7; both SO4 -and HO  are present at pH = 9;

HO  is the predominant radical at pH = 12 [START_REF] Yang | Degradation efficiencies of azo dye Acid Orange 7 by the interaction of heat, UV and anions with common oxidants: persulfate, peroxymonosulfate and hydrogen peroxide [J][END_REF].

The popularity of SO4 -for the purpose of pollutants degradation persuaded us to go for a comprehensive study on application of PMS as a promising source of SO4 -. Following the current investigations, AOPs are shifting towards use of SO4 -due to its unique characteristics.

Hence, PMS is an effective and promising oxidant for producing hydroxyl and sulfate radicals.

PMS-based advanced oxidation processes can open a new perspective of oxidative reactions

to degrade the organic pollutants.

PS-based advanced oxidation processes

Persulfate ion (S2O8 2-, PS) (the most common form is sodium PS (Na2S2O8) and potassium PS (K2S2O8)), usually is colorless or white crystal, and has high stability [START_REF] Liang | Thermally activated PS oxidation of trichloroethylene (TCE) and 1,1,1-trichloroethane (TCA) in aqueous systems and soil slurries[END_REF]. It has symmetrical structure shown in Figure 2-2, the distance of O-O bond is 1.497 Å, and its bond energy is 140 kJ mol -1 [START_REF] Flanagan | The active principle of Caro's acid, HSO5 -: X-ray crystal structure of KHSO5•H 2O [J[END_REF][START_REF] Kolthoff | The Chemistry of PS. I. The kinetics and mechanism of the decomposition of the PS ion in aqueous medium1[END_REF]. Recently, PS has received attention for water remediation due to its high stability, aqueous solubility, high efficiency and relatively low cost [START_REF] Budaev | Degradation of thiocyanate in aqueous solution by persulfate activated ferric ion[END_REF][START_REF] Monteagudo | In situ chemical oxidation of carbamazepine solutions using persulfate simultaneously activated by heat energy, UV light, Fe 2+ , ions, and H2O2 [J[END_REF]. Persulfate is a strong oxidant (E0 (S2O8 2-/SO4 2-) = 2.05 V/ENH) [START_REF] Lau | The aqueous degradation of butylated hydroxyanisole by UV/S2O8 2-: study of reaction mechanisms via dimerization and mineralization[END_REF] and can act directly as an oxidant when activated by heat, UV, electron and transition metals to generate powerful SO4 -radical.

M n+ + S2O8 2-→ M (n+1)+ + SO4 2-+ SO4 - 2-26
Among the different kinds of transition metal ions used for persulfate activation, the most commonly used is iron and iron complexes. These species have been proven to be efficient activators and can be easily injected into soils and groundwater due to their non-toxicity, and low cost [START_REF] Ahmad | Persulfate activation by subsurface minerals [J][END_REF][START_REF] Liang | Persulfate oxidation for in situ remediation of TCE. I. Activated by ferrous ion with and without a persulfate-thiosulfate redox couple[END_REF][START_REF] Zhang | In situ chemical oxidation of aniline by persulfate with iron(II) at ambient temperature [J][END_REF][START_REF] Rastogi | Effect of inorganic, synthetic and naturally occurring chelating agents on Fe(II) mediated advanced oxidation of chlorophenols[END_REF].

In Wu et al. paper [83], Fe(III)/PS system is used to degrade an emerging contaminant 4-tert-butylphenol (4tBP). The results show that, when experiments are conducted at a pH lower than 4 (to guarantee that all the Fe(III) is under dissolved form), 4tBP can be efficiently degraded. The pH has a significant effect on 4tBP degradation due to the speciation of Fe(III).

When using Fe(II)/PS system to degrade other organic pollutants [START_REF] Liu | Oxidative degradation of propachlor by ferrous and copper ion activated persulfate [J][END_REF][START_REF] Pan | Degradation of the UV-filter benzophenone-3 in aqueous solution using persulfate activated by heat, metal ions and light[END_REF], it is shown that the activation of persulfate by Fe(II) generally results in a rapid degradation in the early stage, then follows by a decrease of efficiency due to the rapid depletion of Fe 2+ by the sulfate radicals. This phenomenon can be explained by the eqs 2-27. When the concentration of Fe(II)

is relatively high, it can react with SO4 -to inhibit the reaction. Hence in this Fe(II)/PS system, the initial Fe 2+ dosage is a crucial factor. In order to overcome this disadvantage, sequential and continuous addition of Fe 2+ to activate persulfate could be an alternative strategy. The study by Jiang et al. [START_REF] Jiang | Degradation of bisphenol A in aqueous solution by persulfate activated with ferrous ion[END_REF] indicates that degradation of bisphenol A is enhanced from 49% to 97% using sequential additions of Fe 2+ , and continuous Fe 2+ addition achieved 100% removal using the same amount of total iron and reaction time.

Fe 2+ + SO4 -→Fe 3+ + SO4 2- 2-27
Zero-valent iron (ZVI) can serve as an alternative source of Fe 2+ under oxic or oxidation conditions [START_REF] Zhao | Simultaneous removal of bisphenol A and phosphate in zero-valent iron activated persulfate oxidation process[END_REF] and so for PS activation. Expect for Fe 3+ /Fe 2+ / ZVI, some researches show that iron sulfides such as amorphous FeS, makinawite, troilite (FeS), greigite (Fe3S4), pyrite (FeS2) and marcasite (FeS2) may also activate persulfate by serving as precursors of dissolved or surface Fe 2+ to achieve in situ degradation of organic contaminants just like what Fe 2+ or ZVI did [START_REF] Oh | Degradation of 2,4-dinitrotoluene by persulfate activated with iron sulfides[END_REF][START_REF] Zhang | Efficient pyrite activating persulfate process for degradation of p-chloroaniline in aqueous systems: a mechanistic study[END_REF].

(2) Thermal activation persulfate

Several studies have demonstrated that persulfate can be thermally converted into SO4 - (between the range 30-99℃) as shown in the following eqs 2-28 [START_REF] Liang | Influence of pH on persulfate oxidation of TCE at ambient temperatures[END_REF][START_REF] Frontistis | Removal of cibacron black commercial dye with heat-or iron-activated persulfate: statistical evaluation of key operating parameters on decolorization and degradation by-products[END_REF]:

S2O8 2-+ heat → 2 SO4 - 2-28
The use of heat activation of PS for the degradation of emerging pollutants shows that, under appropriate conditions, most of the pollutants could be efficiently removed [START_REF] Tan | Heat-activated persulfate oxidation of diuron in water [J][END_REF][START_REF] Fan | Kinetic and mechanistic investigations of the degradation of sulfamethazine in heat-activated PS oxidation process [J][END_REF][START_REF] Xu | Mineralization of naphtenic acids with thermally-activated persulfate: The important role of oxygen[END_REF]. Waldemer et al explored the possible application of heat-activated persulfate on chlorinated ethene oxidation [START_REF] Waldemer | Oxidation of chlorinated ethenes by heat-activated persulfate: kinetics and products[END_REF]. The results show that the disappearance of substrates follow the pseudo-first-order kinetics, and rate constants measured from 30 to 70 °C, fit the Arrhenius equation. Concerning the generated products after the reaction, it is observed that dechlorination is one of the most pathways.

Concerning the parameters that affect the efficiency of persulfate thermal activation, it is found that the increase of temperature can accelerate the decomposition of PS to SO4 -.

However, excessive high temperature not facilitate the organic contaminants degradation all the time., In Wang's report, the degradation efficiency of polycyclic aromatic hydrocarbon decreased when the temperature increased over 60°C [START_REF] Wang | Compatibility of surfactants and thermally activated persulfate for enhanced subsurface remediation[END_REF]. However, if the higher temperature enhances the degradation efficiency and solubility of the contaminants into the aqueous phase, temperature optimization is used to minimize the energy input required for achieving desired results. For example, if 100% removal of substrate is accomplished at 60°C and 90°C , 60°C

was chosen for further experiments [START_REF] Costanza | PCE oxidation by sodium persulfate in the presence of solids[END_REF]. Moreover, microwave heating can also have the same better effect on persulfate activation [START_REF] Lee | Decomposition of perfluorooctanoic acid by microwave activated persulfate: Effects of temperature, pH, and chloride ions[END_REF], as microwave heating imparts the required energy more quickly to activate persulfate than conventional method.

(3) UV activated persulfate

As an environmentally friendly and efficient method, UV irradiation activated persulfate to degrade a wide range of organic contaminants was reported by great quantity of previous studies [START_REF] Bu | Degradation of oxcarbazepine by UV-activated persulfate oxidation: kinetics, mechanisms, and pathways[END_REF][START_REF] Dhaka | Aqueous phase degradation of methyl paraben using UV-activated persulfate method[END_REF][START_REF] Hou | Degradation Kinetics and Pathways of Haloacetonitriles by the UV/Persulfate Process[END_REF]. In particular, PS irradiated by UV can generate both active oxidative species SO4 -and HO  to be the oxidants as following eqs 2-29 and 2-30 [START_REF] Tan | Degradation of antipyrine by UV, UV/H2O2 and UV/PS[J][END_REF][START_REF] Luo | Simulation and comparative study on the oxidation kinetics of atrazine by UV/H2O2, UV/HSO5and UV/S2O8 2[END_REF]:

S2O8 2-+hv → 2 SO4 - 2-29 SO4 -+ H2O → HSO4 -+ HO • 2-30
The degradation kinetics and mechanism of the antiepileptic drug oxcarbazepine (OXC) by UV-activated persulfate oxidation are investigated. The results show that UV/persulfate process appeared to be more effective in degrading OXC than UV or PS alone, and its degradation exhibited a pseudo-first order kinetics pattern [START_REF] Bu | Degradation of oxcarbazepine by UV-activated persulfate oxidation: kinetics, mechanisms, and pathways[END_REF]. SO4 -and HO  are identified to be responsible for substrate degradation and SO4 -made the predominant contribution in the process. For this UV/persulfate system, the most common wavelengths of the UV light are between 200 nm and 300 nm, as the maximum absorption of many organic contaminants occurs within this range [START_REF] Luo | Simulation and comparative study on the oxidation kinetics of atrazine by UV/H2O2, UV/HSO5and UV/S2O8 2[END_REF][START_REF] Ao | Degradation of sulfamethoxazole by medium pressure UV and oxidants: Peroxymonosulfate, persulfate, and hydrogen peroxide[END_REF]. Quantum yield is an important factor for characterizing the activation of PS, and UV wavelength has significant influence on the quantum yields. It was found that the quantum yields of sulfate radicals decreased with increase of UV wavelength in the range of 248 to 351 nm [START_REF] Herrmann | On the Photolysis of Simple Anions and Neutral Molecules as Sources of SOxand Cl in Aqueous Solution[END_REF]. The maximal quantum yield is about 1.4 at the wavelength of 248 and 253.7 nm. In general, 254 nm is most used as the irradiation wavelength for persulfate, not only because which is in the energy absorbance of the ultraviolet range and can activate the PS to generate SO4 -with a quantum yield up to 0.7 mol E/s [START_REF] Mark | The photolysis of potassium peroxodisulphate in aqueous solution in the presence of tert-butanol: a simple acidometer for 254 nm radiation [J][END_REF], but also it can decrease reaction time requirements compared to other wavelengths.

A comparative study using 254 nm and 365 nm UV activated persulfate for polyvinyl alcohol destruction found that 100% removal was achieved in less than 5 min at 254 nm but only 93% removal in 30 min was observed using 365 nm [START_REF] Lin | UV/S 2 O 8 2-, process for degrading polyvinyl alcohol in aqueous solutions[END_REF]. In addition, Gamma ray has been also demonstrated to be capable of activating the PS [START_REF] Devi | In-situ chemical oxidation: Principle and applications of peroxide and persulfate treatments in wastewater systems[END_REF][START_REF] Matzek | Activated PS for organic chemical degradation: a review[END_REF].

(4) Electrochemically-activated persulfate Electrochemical activation of PS is a promising green technology with electrons acting as the reactant, which has drawn increasing attention in recent years [START_REF] Bu | Removal of 2-MIB and geosmin by electrogenerated persulfate: Performance, mechanism and pathways[END_REF][START_REF] Farhat | Removal of Persistent Organic Contaminants by Electrochemically Activated Sulfate[END_REF]. Electrochemical reactions generate the sulfate radical at the cathode (eqs 2-31) following the same mechanism as the one electron transfer redox reaction for iron-activated persulfate [START_REF] Zhao | Enhanced oxidation of 4-chlorophenol using sulfate radicals generated from zero-valent iron and peroxydisulfate at ambient temperature[END_REF] S2O8 2-+ e -→ SO4 -+ SO4 Furthermore, electrochemical regeneration of iron for persulfate activation can reduce the total iron used in these systems by using the current to control iron release, eliminating the need for further iron dosing [START_REF] Lin | Degradation of bisphenol A in aqueous solution by a novel electro/Fe 3+ /peroxydisulfate process[END_REF]. Expect for iron, electrochemical persulfate activation also can be accomplished by using platinum anodes and cathodes which can eliminate the need for iron addition [START_REF] Jhou | Mineralization of dinitrotoluenes in industrial wastewater by electro-activated persulfate oxidation[END_REF]. Electrochemical techniques have proven to fully mineralize recalcitrant compounds not degraded via activated-persulfate alone [START_REF] Carter | Oxidative destruction of perfluorooctane sulfonate using boron-doped diamond film electrodes[END_REF]. Hence, combining electrochemistry and electrochemically-activated persulfate may achieve interesting results compared to either technique alone, thus reducing persulfate and/or electricity usage.

However, more studies should be conducted to find the best suited electrode materials for activating persulfate, the relationship between current density and activation. Moreover, studies are needed to determine which electrode materials, the synergistic effects on degradation and mineralization of various classes of organics.

During the activation process of PS, PS itself can be decomposed. The decomposition of PS varied with the activation methods. Decomposition of PS is affected not only by activation methods but also by the experimental conditions. It is noted that high decomposition efficiency of PS or PMS does not mean the good performance for removing the organic compounds. Therefore, the decomposition efficiency of PS cannot be used to reflect the removal capacity of the selected process. Instead, the reaction stoichiometry efficiency (the ratio of degraded organic compounds concentration to consumed oxidant concentration) could be used to evaluate the removal capacity of the selected process.

The activated PS has the potential for the degradation of emerging contaminants in soil, groundwater, water and wastewater. Among the various activation methods, ultraviolet and metal oxide activation are the most investigated methods according to the published references. In summary, UV/PS process is potential to application for removing the recalcitrant compounds from wastewater, but the effect of complex wastewater compositions should be pay attention too. For metal activated PS, Fe(Ⅱ)-activated persulfate could be a promising process in terms of both economic and water safety. Although many methods can be used to activate PS and PMS, there is still a long way to go for the practical application.

Sulfite-based advanced oxidation processes

Sulfite (SO3 2-) or bisulfite (HSO3 -) are the most common S(IV) species. S(IV) is important for wine making, fruit preservation, or as dehydrated food additive and textural altering agents

HSO3 -+ Fe 2+ → FeHSO3 + 2-35 4FeHSO3 + + O2 → 4FeSO3 + + 2H2O 2-36 FeSO3 + → Fe 2+ + SO3 - 2-37
As the sulfite radical (SO3 -) is generated during this oxidation process, the autoxidation of sulfur(IV) in aqueous solution is involved in a chain reaction propagated by oxysulfur radicals

[125] The species involved are the sulfite radical (SO3 -), the peroxomonosulfate radical (SO5 -), and the sulfate radical (SO4 -) as suggested by Hayon et al. [126]. The reaction chain is initiated by the formation of SO3 -, which can be generated from eqs 2-35 to 2-37, In the presence of oxygen, SO3 -is rapidly converted to SO5 -, which acts as the main chain carrier, then undergoes further reaction to form SO4 -.

Nowadays, sulfite, a typical reducing agent, has been widely used for remediation of contaminated sites in many emergency cases. Moreover, the significant role of sulfite as an efficient source of active radicals via the reaction of an iron-sulfite-oxygen complex has been reported. Chen et al. reported this new process for dyes degradation with SO4 -using an Fe(II)/sulfite system. Main results show that efficient degradation of dyes could be achieved using this novel system in acidic environment in comparison with the Fe(II)/persulfate and Fenton (Fe(II)/H2O2) systems [START_REF] Long | Decolorization of Orange II in Aqueous Solution by an Fe(II)/sulfite System: Replacement of Persulfate[END_REF]. In Zhang et al's report [START_REF] Zhang | Enhanced Decolorization of Orange II Solutions by the Fe(II)-Sulfite System under Xenon Lamp Irradiation[END_REF], 84% of Orange II (10 mg L -1 ) can be efficiently decolorized within 60 min in an Fe(II)/sulfite-based photochemical system using a xenon lamp (350 W), whereas less efficiency (15%) was achieved without irradiation. Application of industrially available sulfite as a new source of sulfate radicals, instead of using expensive persulfate reagents, seems very promising in the practice of advanced oxidation technologies. Chromium, a common heavy-metal contaminant, is known as a hazardous material. It normally occurs as hexavalent chromium (Cr(VI)) and trivalent chromium (Cr(III)) [1]. As Cr(VI) is more toxic than Cr(III), reduction of Cr(VI) to Cr(III) is the principal strategy for pollution control for industrial emissions containing Cr(VI) [2]. Because of the high standard reduction potential of Cr(VI)/Cr(III) at acidic pH (e.g., E 0 : Cr2O7 2-/Cr 3+ = 1.36 V) [3], Cr(VI) can be easily reduced by chemical reductants [4,5] and biological processes [6]. Among its known reductants, ferrous iron (Fe(II), E 0 : Fe 3+ /Fe 2+ = 0.771 V) [3] has been extensively used over the past several decades because it is abundant, reliable, inexpensive, and efficient [7,8].

Another advantage of using Fe(II) as reductant in wastewater is the simple separation of reduced Cr(III) via Cr(OH)3 formation and coprecipitation with Fe(OH)3 at high pH.

Industrial wastewaters usually contain more than one type of contaminant. Even effluents from electroplating and textile industries contain heavy metals and organic additives [9,10].

Therefore, heavy metals and organic contaminants must be removed to meet the requirement of discharge standards. Dual functions have been developed for simultaneous removal of heavy metals and organic contaminants. Methods such as biodegradation, biosorption, adsorption, and nanotechnology using Pd(II) and Cr(VI) for removing organic pollutants such as benzoate, as well as phenol catabolism, have received much attention [11][12][13]. Kaur and Crimi designed a system using in situ chemical oxidation and persulfate to treat soils and groundwater contaminated with organic and metal contaminants [14]; however, its degradation processes take much time and involve many complex technical procedures to achieve the goal. Semiconductor photocatalytic systems (e.g., TiO2) can produce reductive and oxidative species (e.g., electron-hole pair) simultaneously [15]. The photo-Fenton-like process induced by the Fe(III)-OH complex [16] produces reductive Fe(II) and oxidative HO  radicals. These photochemical systems have been applied in simultaneous reduction of Cr(VI) and oxidation of organic contaminants (e.g., dye and surfactant); however, they are more expensive and complicated.

Studies have also used sulfur species such as sulfides (hydrogen sulfide [17] and calcium polysulfide [18]) and sulfite [19] for Cr(VI) reduction. These works focused on the kinetics of Cr(VI) reduction but did not probe the mechanism of transformation of sulfur species.

Cr(VI)-catalyzed autoxidation of S(IV) [20,21] had been investigated during 1960s and 1990s.

Formation of oxysulfur radicals (mainly SO3 -, SO4 -, and SO5 -) in this reaction has been confirmed; however, the potential application of such oxysulfur radicals in the oxidation of organic or inorganic contaminants, except in the Fe(III)/Fe(II)-S(IV)-O2 system, has been very limited [22,23]. Recently, we showed that the rapid decolorization of azo dyes by the Fe(III)/Fe(II)-S(IV)-O2 system with or without UV-vis irradiation [24][25][26] is a promising alternative to an advanced oxidation process using persulfate [27]. The possibility of oxidation of organic contaminants by a S(IV)-O2 system using Cr(VI) induction and Cr(VI) reduction has not yet been explored.

The aim of this work was to propose the simultaneous reduction of Cr(VI) and oxidation of organic contaminants by sulfite ions in the presence of oxygen. The Cr(VI)-S(IV)-O2 system reduces Cr(VI) during S(IV) autoxidation, which means that it can utilize inorganic waste Cr(VI) to degrade organic waste. A rapid reaction using the azo dye acid orange 7 (AO7, the organic substrate) that is comparable to Cr(VI) reduction was used in oxidative decolorization at various pH values and Cr(VI)/S(IV) molar ratios. We examined the effect of oxygen and different kinds of radical scavengers used to quench radicals generated in this system. The mechanism of the system and the contribution of reactive oxygen species and reactive sulfur species were preliminarily probed. Furthermore, aniline, phenol, bisphenol A and their substituted analogs were selected to testify the possible application of this system. This work provides a novel and low-cost system for waste treatment using waste Cr(VI) to achieve simultaneous reduction of Cr(VI) and oxidation of organic contaminants.

Material and methods

Chemicals.

A Cr(VI) solution was prepared by dissolving analytical-grade (AR) potassium dichromate (K2Cr2O7). AO7, 1,5-diphenylc arbazide used to detect the concentration of Cr(VI), the radical scavengers tert-butyl alcohol (TBA) and aniline, as well as NaOH (AR) and H2SO4

(AR) which were used to adjust the pH of the solutions, were obtained from Sinopharm Chemical Reagent Co., Ltd. Na2SO3 (AR) (Shanghai Zhanyun Chemical Co., Ltd.) was prepared just prior to measurements. Analytical-grade o-Nitroaniline, p-Nitroaniline, p-Chloroaniline, Sulfonamides, Phenol, Bisphenol A, 2-Bromophenol, Paracetamol, p-aminophenol, Catechol, o-Nitrophenol, 2,6-Dinitrophenol, which were used to testify the possible application of this system were also purchased from Sinopharm Chemical Reagent Co., Ltd. Acetonitrile and methanol were HPLC grade and purchased from Fisher Corporation.

All chemicals were used without further purification. Ultrapure water with 18.2 MΩ cm resistivity used in this work was obtained through a water purification system (Liyuan Electric Instrument Co., Beijing, China).

Reaction procedures.

All experiments were performed by using a 200 mL beaker at about 20 ± 2 C and under normal laboratory lighting. Predetermined amounts of AO7 and Na2SO3 were added to the beaker. The pH of reaction solutions thus prepared was adjusted to the desired values with dilute NaOH and H2SO4. Each solution was constantly stirred with a Polytetrafluoroetylene (PTFE)-coated magnetic stirrer, and each run was performed by careful addition of the desired amount of Cr(VI) that was prepared beforehand. The pH was not controlled during the reaction. Samples were withdrawn at specific time intervals and analyzed immediately in order to avoid further reaction. The initial concentration of the substrates in all batch experiments was 0.029 mM. In the radical-scavenging experiments, specific concentrations of TBA or aniline, scavengers for HO • or oxysulfur radicals, respectively, were added to the reaction solutions (the procedure for preparation of their solutions is similar to that for substrates and Na2SO3). The concentrations of TBA and aniline in the solutions were 11.6 and 2.9 mM, respectively. The effect of oxygen was also investigated. Nitrogen introduced through a gas-purging tube was used to exclude oxygen from the solution. To maintain constant conditions, the total flow rate of nitrogen was maintained at 1 L min -1 . All experiments were carried out in parallel. All relative errors were <5%.

Analysis.

An optical-fiber spectrometer (Shimadzu, Japan) was used to measure AO7 concentrations by absorbance detection at 485 nm. The Cr(VI) concentration was determined spectrophotometrically at 540 nm by using 1,5-diphenylcarbazide, and the calibration of Cr(VI) was shown in Figure 3-1. In addition, spectra of the remaining AO7 were obtained on a UV-1601 spectrophotometer (Shimadzu, Japan) using a cuvette with 1 cm width. A LC-10A HPLC system (Shimadzu, Japan) equipped with a SIL-10A auto sampler, an SPD-10A

UV detector and a C18 reverse-phase column (4.6 × 250 mm, 5 μm, Dalian Elite Analytical Instruments Co., LTD., PRC) was used to monitor residual aniline, phenol, bisphenol A etc.

The injected volume was 100μL ,the UV detection and the corresponding mobile phase with a HCrO4 -ion is the only predominant species of Cr(VI) at pH 1 -6, while the fraction of HCrO4 -drops sharply when increasing pH from 6 to 9. This figure was created by the special software MEDUSA. Besides the distributions of Cr(VI) and S(IV) species, reactions in eqs (3-3) and (3)(4) are also pH-dependent; the latter have higher rates at acidic pH. These are the original source of SO3 -, which requires sufficient amounts of H + to form the stable Cr(III)-S(VI) complex. The rate of SO3 -production may determine the subsequent production rates of the oxysulfur radicals SO4 -and SO5 -in the presence of a fixed concentration of S(IV) and atmospheric oxygen. As Brandt and Elding [21] indicated, the overall rate of this reaction decreases as the pH increases within the range of 4-6. In the present work, the overall efficiencies of AO7 decolorization and Cr(VI) reduction at pH 3.0 were higher than those at other pH values.

Consequently, pH 3.0 was deemed as the optimal pH and was used in subsequent experiments. HSO3 -ion is the only predominant species of S(IV) at pH 3 -6, while the fraction of HSO3 -drops sharply when increasing pH from 1 to 3 or 6 to 9.

Effect of the Cr(VI)/S(IV) molar ratio.

As shown in the previous section, the initial pH of the reaction mixture is of great importance to AO7 decolorization and Cr(VI) reduction in the aqueous Cr(VI)-S(IV)-O2 system. For the sake of practicality, the Cr(VI)/S(IV) molar ratio needed to ensure complete reaction at the optimum pHinit was determined. The efficiency of AO7 decolorization at various Cr(VI)/S(IV) molar ratios at initial AO7 concentration of 0.029 mM and at pH 3.0 was determined. Results

in Figure 3 The aforementioned results illustrate that an increase in S(IV) dosage did not always enhance AO7 decolorization; excessively high S(IV) concentrations (such as 1.0 mM) could retard the subsequent reaction in the system. Several scholars [28,29] have confirmed that at high [M(n)]/[S(IV)] and low [S(IV)]/[O2] ratios (M(n): heavy-metal ion), the system undergoes oxidation. Otherwise, the system tends to predominantly undergo reduction. Therefore, elevated concentrations of S(IV), a reducing species, may prevent the system from undergoing oxidation. However, an increase in S(IV) concentration could facilitate AO7 decolorization because it generates oxysulfur radicals. Hence, the balance between the reducing nature of S(IV) and its catalytic character is critical. Considering the factors above, we deemed 0.1:0.55 mM as the optimum Cr(VI)/S(IV) molar ratio for this system.

Effect of oxygen and radical scavengers.

The effect of oxygen on AO7 decolorization in the Cr(VI)-S(IV)-O2 system was investigated (results are presented in Figure 3-10). When the experiment was conducted in the presence of air, the decolorization efficiency was nearly 57.7%. The efficiency dropped to 14.6% upon exclusion of oxygen from the solution by purging with nitrogen. This result implies that oxygen plays an important role in AO7 decolorization in the system. Furthermore, as long as there is dissolved oxygen in the solution, SO3 -could react rapidly with oxygen, forming SO5 -and SO4 -(eqs ((3-5)-(3-9)), [26,28] which can oxidatively decolorize AO7. In addition, reaction of SO4 -with H2O can produce HO • (reaction (3-12)), which is alternative for AO7 decolorization either [30]. To shed light onto the reaction mechanism, we conducted quenching studies to identify the activity of the added radicals, TBA and aniline. In our previous studies [31], TBA could completely inhibit the reaction of AO7 with HO • when [TBA]0 = 400[AO7]0. HO  is an extremely potent oxidizing agent that unselectively oxidizes various kinds of organic pollutants [32,33]. However as shown in Figure 3-10, inhibition with TBA resulted in only a slight decrease in decolorization rate, indicating that HO  did not play a significant role in AO7 decolorization in the Cr(VI)-S(IV)-O2 system. The pseudo-first-order kinetic law was applied to analyze the decolorization process of AO7 (examples of the regression analysis of the pseudo-first-order kinetic curves were shown in Figure 3 -11). The pseudo-first-order rate (kobs) in Table 3-1 suggest that the contribution of HO  to AO7 decolorization was only ~21.7%. Aniline may be an appropriate scavenger for SO4 -, as the rate constant (k) with aniline and SO4 -is on the order of 10 9 L mol -1 s -1 (data estimated in our other work not shown). The rate constant for the reaction between aniline and HO  (k = 1.7 × 10 10 L mol -1 s -1 ) [34] is nearly three orders of magnitude larger than those for the reactions between aniline and SO5 -(k = 3.0 × 10 6 L mol -1 s -1 ) and between aniline and SO3 -(k < 1.0 × 10 6 L mol -1 s -1 ) [35]. However, when the concentration of aniline added to the reaction was high enough, almost all of the oxysulfur radicals (SO4 -and SO5 -) and HO  could be quenched completely [31]. When [aniline]0 = 100[AO7]0, <5% of AO7 was decolorized, confirming that oxysulfur radicals are the main active species in the Cr(VI)-S(IV)-O2 system under acidic conditions.

SO4 -+ H2O → SO4 2-+ HO  + H + 3-12
The contribution of oxysulfur radicals (SO4 -and SO5 -) to AO7 decolorization was nearly 72.5% (Table 3-1), which confirms that oxysulfur radicals were the oxidants that mainly led to decolorization. a SOx •-refers to oxysulfur radicals (mainly SO4 •-and SO5 •-). UV-vis absorption spectra of the remaining AO7 during the reaction (Figure. [3][4][5][6][7][8][9][10][11][12] indicate a decline in absorbance at 485 nm, which suggests that the Cr(VI)-S(IV)-O2 system could effectively destroy the azo bond, the AO7 chromophore, resulting in decolorization. However, the efficiency of TOC removal, which was the general index used to investigate the mineralization of organic pollutants, did not decline even after 48 h of reaction. The efficiency of TOC removal in the Fe(II)-S(IV)-O2 system in our previous studies [24,25], is very similar to that in this work (data not shown). Therefore, the Cr(VI)-S(IV)-O2 system is more reliable for decolorization than for mineralization.

Appliance of Cr(VI)-S(IV)-O2 system for degrading other organic compounds

Cr(VI)-S(IV)-O2 system not only has strong decolorization effect on azo dye, similarly, rapid degradation of other organic pollutants such as aniline, phenol, bisphenol A and their substituted analogs (more than 10 derivatives) within 60 min can also be achieved. As shown in Table 3-2, the experiments were conducted at specific Cr(VI)/S(IV) molar ratios (0.1: 0.55 mM) at initial substrates concentration of 0.029 mM and at pH 3.0. After 30 min reaction, all the pollutants selected in this study underwent different degree of degradation, and almost no further degradation occurred after 60 min reaction. However, the degradation rate of organic pollutants is quite different, and the reason may be explained by the existence of different substituted groups. In general, the degradation efficiency of phenol and its derivatives are much higher than aniline and its derivatives in the same condition. However, referring to the influence of functional group, for both aromatic amines and phenols, the same substituted group has the same trend on degradation. As shown in In addition, the results in Table 3-2 also show that the degradation efficiency exhibited a significant difference for aniline, phenol and their derivatives with the same substituents but at different substituted positions. Electronic parameter such as Hammett σ constants, simply obtained from the ionization of organic acids in solution, can frequently predict successfully equilibrium and rate for a variety of families of reactions in solution [36]. The combination of electronic, steric, hydrophobic/hydrophilic, and hydrogen-bonding parameters has been used

to derive quantitative structure-activity relationships (QSAR) for a host of reactions of organic compounds in aqueous solutions [37]. In this study, the relationship between the Hammett σ values for each compound and the pseudo-first-order rate (kobs1) has been investigated. The kobs1 were listed for the substituents with Hammett σ values [38][39] in Table 3-3, which showed that the reaction rate dramatically increased with abundant electron.

Furthermore, the σ corresponds to the effect that a substituted group would have upon a benzene ring if it was substituted at different position. shows that the calculated t value is 2.70, which is above the critical value of t, 2.571 (N= 7, significance level = 0.05 (two tails)).

Conclusions

The Cr(VI)-S(IV)-O2 system is an efficient system for rapid and simultaneous reduction of Cr(VI) and oxidative degradation of organic pollutants in aqueous solution with dioxygen under acidic conditions. We have considered effects such as pH, concentration of added Cr(VI)

and S(IV), as well as the presence of O2 or N2 and various radical scavengers. Our results

show that the reaction of Cr(VI) and degradation of organic pollutants depended heavily on the pH and initial concentrations of Cr(VI) and S(IV). In addition, oxysulfur radicals were the dominant oxidants for AO7 decolorization. Thus, our Cr(VI)-S(IV)-O2 system may be a new candidate for use in the efficient treatment of waste using waste chromium via simultaneous reduction of Cr(VI) and oxidation of organic contaminants.

Introduction

Advanced oxidation processes based on sulfate radical (SR-AOPs) have emerged as a promising method in the field of oxidative decontamination of polluted water and soil [1][2][3].

Sulfate radical (SO4 -), a strong one-electron oxidant, has relatively high standard redox potential (E 0 = 2.6 V vs NHE) with an oxidation potential comparable to or even higher than that of hydroxyl radical [4]. Moreover, SO4 -can react via electron transfer, by addition to C-C double bonds and H-abstraction [5,6], thus, it is able to oxidize a large number of pollutants such as phenol derivatives and aniline in water [7][8][9]. SO4 -can be generated in homogeneous or heterogeneous systems via photolysis, thermolysis and radiolysis [10,11] or via transition metal activation of persulfate (S2O8 2-, PS) [7,[12][13][14][15] and peroxymonosulfate (SO5 2-, PMS) [16,17].

In fact, PMS can be activated by various transition metals such as Fe, Mn, Ni and Co in the homogenous systems [1,18]. Among them, Co and Fe are the most commonly used metal to promote radical formation due to their occurrence in natural media and low cost. Huang and

Huang investigated the ability of Co(II) and PMS system to degrade Bisphenol A at pH 7, and achieved an efficient detoxification and mineralization method [19]. A process based on the sulfate radicals generation through iron (Fe(II), Fe(III)) activation of PMS or PS was studied for polychlorinated biphenyls degradation in aqueous system [20]. The high oxidation efficiency and slow rate of consumption of the oxidants make metal-mediated activation system a feasible strategy for degradation of recalcitrant organic compounds. Furthermore, PMS activation using cobalt oxide or cobalt-metal oxide as heterogeneous catalysts also gains significant relevance in water treatment applications [21]. The cobalt oxides such as CoO, CoO2, CoO(OH), Co2O3 and Co3O4, Fe-Co mixed oxide nanocatalysts, cobalt oxide supported on MgO (Co/MgO), on TiO2 (Co/TiO2) and Co3O4/TiO2 and combined with other metals were used as efficient heterogeneous catalysts for activation of PMS [22][23][24][25]. The most advantage of heterogeneous catalyst is that solid particles can be easily removed from liquid phase and, in some cases, reused.

Recently, sulfite ions were found to react with transition metals such as Fe(II), Fe(III) and Cr(VI) to generate SO4 -and application for the azo dyes and amine compounds decontamination was tested [17,[26][27][28][29][30].

In our previous work, we reported some novel AOPs using Fe(II)-sulfite, Fe(III)-sulfite, photo-Fe(II)-sulfite system able to produce oxysulfur radicals (including SO3 -, SO4 -and SO5 -). Combined with the work conducted by other researchers [31][32][33], the basic chain oxidation mechanisms of oxysulfur radicals generation has been investigated.

In this work, Co(II)-sulfite ions (S(IV)) system has been investigated to promote the paracetamol degradation in water. Paracetamol (PARA) is a widely used analgesic and antipyretic drug and an important material for the manufacturing of azo dyes. PARA was chosen as a target contaminant in this work due to its presence in the environment from several emissions from manufacturing facilities, consumer use and disposal, and hospital waste [34,35]. The effect of pH, initial PARA concentration, Co(II)/S(IV) molar ratio, the presence of oxygen were investigated. Moreover, the activation mechanism and contribution of reactive oxygen and sulfur species were elucidated by using different kinds of radical scavengers and transient absorption spectroscopy.

Materials and methods

Degradation experiments

All experiments were conducted in a 250 mL open cylindrical reactor cooled by an external jacket water circulation at a constant temperature of 25°C. Appropriate amounts of the PARA, scavengers and Na2SO3 were mixed in the solution and the pH was adjusted using a pHS-3C pH meter by adding dilute NaOH and H2SO4 until desired value. Each solution was constantly stirred with a Polytetrafluoroethylene (PTFE)-coated magnetic stirrer and purged with air with a fixed flow of 0.8 L min -1 .

Each experiment was initialed spiking with Co(II) or CoO into the solution. Samples were withdrawn at fixed interval times and analyzed immediately to determine the remaining concentration of PARA and Co(II). For radical-scavenging experiments, specific concentrations of TBA or EtOH were added to the solutions before Co(II) addition. In order to assess the role of oxygen during Co(II)-sulfite solution were purged by bubbling N2 ( 99.99%) or O2 (99.99%) for 30 min before and throughout the experiment. A dissolved oxygen (DO) meter (8403, AZ Instrument Co. Ltd.) was used to determine the oxygen concentration in solution. During anoxic reaction, the DO was 0.01 mg L -1 while for O2 saturated solution a concentration of 20 mg L -1 was determined. All experiments were carried at least two times.

Chemical analysis

The concentration of PARA was determined using a high-performance liquid chromatography (HPLC) Shimadzu LC-10A system equipped with UV-vis detector (SPD-10AV; Shimadzu) and an ODS-C18 column (25 cm × 4.6 mm, 5 μm; Shimadzu, Kyoto, Japan). The separation was carried out using methanol: water (25:75 v/v) as isocratic mobile phase at a flow rate of 1.0 mL min -1 . The detector was set at 241 nm. An optical-fiber coupled to a spectrophotometer (UV-1601 Shimadzu, Japan) was used to scan the UV-vis absorption spectra of Co(II)-SO3 2-complex. The concentration of Co(II) in solution was determined by

Methyl isobutyl ketone and Ammonium thiocyanate , and the method description is like this:

The calibration of Cobalt(II) was measured by preparing a series of standard solutions, firstly ,5 mL cobalt(II) and 1 mL NH4SCN(566g/L) was added to a 10 mL centrifuge tube and hand shaken for 1 min, then another 4 mL pure MIK was added to the mixture, and ultrasonic shaking for 1min and equilibrating until stratication occurred. Using the UV-Vis spectrophotometer for the absorbance at 620 nm for the upper layer complex which containing Cobalt(II). The calibration curve is presented in Figure 4-1 [36]. The calibration equation is:

[Co 2+ ] (M) = (6.7710 -4  Abs620 nm) -7.0210 -6 For SO4 -reactivity a laser flash photolysis (LFP) system was used with 266 nm excitation following the procedure reported by Wu at al. [15]. Analysis of transient decay and second order rate constant determination (using PARA concentrations from 0 to 0.1 mM) were determined at 470 nm corresponding to the maximum absorption of SO4 -. The second order rate constants were determined at pH 5.0 and 11.0 corresponding to the molecular and deprotonated forms respectively of PARA (pKa = 9.5).

Results and discussion

Control experiments

A typical degradation of Paracetamol by three different kinds of system (only Co(II), only sulfite(S(IV), Co(II)+S(IV)) is presented in Figure .4-2. The result showed that when the experiment conducted in the presence of S(IV) without any catalyst, negligible change in paracetamol concentration was observed. A similar trend was also achieved in the experiment by only using Co(II) even after being centrifuged for 15 min at 9000 rpm. There was no precipitation occurred, which illustrated that there is no adsorption in this system and neither S(IV) nor Co(II) added to the reagent had an effect on Paracetamol. However, when both S(IV)

and Co(II) were in solution, Paracetamol can be degraded to more than 80% in 10 min. These results suggest that Paracetamol degradation depends heavily on the reaction between S(IV)

and Co(II), it proceeded only when S(IV) and Co(II) were present.

Effect of initial pH on PARA degradation

The effect of the initial pH on the Co(II)-S(IV) system oxidation of PARA was investigated the formation of insoluble Cobalt-hydroxide (Co(OH)2) complex that is expected to precipitate in solution [37].

Moreover, it has been reported that deprotonated complexes (Co(II)-OH), which is more reactive toward SO3 2-compared to HSO3 -would be formed at alkaline pH values (eqs 4-1) leading to the formation of a Co(II)-SO3 complex [27]. Co(II)-SO3 could be oxidized to Co(III)-SO3 + complex in the presence of dissolved oxygen as reported in different works (eqs 4-2 and eqs 4-3) [38]. SO3 -could be generated during the redox reaction between the Co(III)-SO3 + complex and SO3 2-(eqs 4-4).

UV-vis absorption spectra of Co(II), S(IV), Co(II) and S(IV) during the reaction (Figure. were acquired to investigate the complexation between Co(II) and S(IV). The result showed that: only Co(II) or S(IV) at pH 9.0 did not show absorption in the range 250-600 nm.

However, when S(IV) was added to the solution containing 0.1 mM Co(II) an absorption band absorbing up to ~ 600 nm is present. The presence of this new absorption band demonstrates that the Co(II)-S(IV) complex could be formed and decreasing in the absorption during time proved the reaction between Co(II) and S(IV), and corresponding depletion of S(IV). The pH represents also a key factor influencing the SO3 reactivity in water. In fact, the primary step is the oxygen-mediated oxidation of SO3 suggest that in the presence of enough sulfite in solution, the cobalt recycle could achieved very fast leading to the generation of SO3 at the same time. Even the addition of Co(II) didn't play a relevant role on the paracetamol degradation, but for the sake of feasibility and accuracy, 0.1 mM Co(II) was employed during all experiments, as when the experiments were conducted under low concentration of Co(II) such as 0.05 mM, the repeatability could not be ensured every time. The dependence of the PARA degradation on the initial concentration of S(IV) was also investigated (Figure 4567). An increase in S(IV) concentration (between 0.5 to 1 mM) accelerate the PARA degradation being , through Co(II) reactivity, a source of SO3 -.

( ) ( )

Co II H O Co II OH    alkaline pH 4-1 2 3 3 2 ( ) ( ) Co II OH SO Co II SO H O       4-2 3 2 3 ( ) ( ) Co II SO O Co III SO      4-3

S(IV) and Co(II) effect on the PARA degradation

However, at higher sulfite ions concentrations (2 mM) there is no more enhancement of PARA degradation and this effect could be attributed to the reactivity competition of SO4 -/SO5 -between PARA and SO3 2-(eqs 4-8 and eqs 4-10). Hence, the balance between S(IV) concentration and its catalytic character is a key factor for the oxidative system efficiency.

In all these experiments (Figure .4-5 and 4-7) it is impossible to obtain more than 90% of PARA degradation. This observation is certainly due to a competition reaction of the sulfate radicals on PARA, the excess of sulfite and the degradation products of PARA. After a short period of time sulfate radicals react no more on paracetamol due to its low concentration.

Effect of PARA initial concentration on the PARA degradation

To investigate the effect of initial PARA concentration on PARA degradation in the Co(II)-S(IV) system, experiments were conducted at different initial PARA concentrations

Effect of O2

The effect of oxygen concentration on the PARA degradation using 0.1 mM of Co(II) and 1 mM of S(IV) was investigated comparing results from aerated and nitrogen-purged solutions.

When experiment was conducted in aerated solution (under air bubbling), a PARA degradation plateau was reached after first 10 min corresponding to ~82% of degradation . However, under nitrogen-saturated solution, the efficiency dropped to less than 8% confirming that oxygen is a crucial parameter. Oxygen strongly favors the degradation of pollutant through SO3 -oxidation into SO5 -that undergo further reaction to form SO4 -in the solution (eqs 4-5 -eqs 4-7).

Radical species involvement

The second order rate constants of sulfate radical with molecular and deprotonated PARA (at pH 9.0, pH used in our experiments, about 25 % of PARA is under deprotonated form) were determined from the linear fit of pseudo-first order decay monitored at 470 nm (corresponding to the maximum absorption of sulfate radical) vs concentration of PARA in solution . The second order rate constant was estimated to be 1.33 ± 0.79 × 10 9 M -1 s -1

(at pH 5) and 6.14 ± 0.99 × 10 8 M -1 s -1 (at pH = 11.0) corresponding respectively to the molecular and deprotonated form, on the phenol group, of PARA.

To shed light onto radical mechanism involved in such system, EtOH and TBA were used as radical scavengers. This competition kinetic approach is based on the different second-order rate constant with HO  and SO4 -( , TBA HO k  = 6.0 × 10 8 M -1 s -1 [44] which is nearly three orders of magnitude higher than presence of TBA (up to 6 mM) no effect was observed on the PARA degradation [Figure. [4][5][6][7][8][9][10][11] indicating that hydroxyl radical are not generated using Co(II) 0.1 mM and SO3 2-1 mM at pH 9.0. Indeed, if hydroxyl radical were produced in the system, addition of 1 mM TBA should be scavenged about 98 % HO  reducing drastically the PARA degradation. However, when EtOH was used as radical scavenger in solution (Figure. [4][5][6][7][8][9][10][11][12] PARA degradation rate and efficiency (after 30 min) were modified. In the presence of 500 mM of EtOH, a complete inhibition of PARA degradation should be expected on the basis of second order rate constant reported before. But, PARA is still degraded with ~ 29% of disappearance after 30 min. Such trend could be explained considering the possible involvement of SO5 -during PARA degradation. In fact, EtOH is not able to scavenge strongly SO5 -due to the very low rate constant ( 5 , EtOH SO k  < 10 3 M -1 s -1 [46]). Such results suggest that SO4 -represent the main radical leading to the degradation, but also possible implication of SO5 -during degradation of PARA is possible.

Heterogeneous reaction between CoO and SO3 2-

Cobalt oxide such as CoO, CoO2, CoO(OH), Co2O3 and Co3O4 were usually used to react with PMS to oxidize different kinds of organic pollutants [24]. Experiments to prove the efficiency of sulfite activation by CoO were conducted at pH 9.0 with the paracetamol concentration of 10 μM and SO3 2-at 1.0 mM. As shown in Figure . 4-13, when 0.1 mM and 1.0 mM CoO were added to the solution, the degradation efficiency were 32.9% and 71.7%, after 30 min respectively, which indicates that CoO exhibit a good catalytic activity. One of heterogeneous catalysts advantages belong to their stability and their reusability as catalyst.

However, oxide can easily agglomerate during catalytic reaction, resulting in the reduction of catalytic performance [47], also cobalt ion leaching and dissolving problem can cause the same potential environmental and health problem as homogeneous catalysts do. As reported in

Conclusions

Our results showed the high efficiency of this Co(II)-S(IV) system using PARA as organic pollutant model in aqueous solution at alkaline pH. In this system, the degradation of paracetamol depend on initial concentrations of S(IV) and strongly the pH. In fact, it is clear that dissolved oxygen plays a crucial role allowing the oxysulfure radicals oxidation to initiate the reaction. The results of radical scavenger experiments demonstrate that SO4 -and also SO5 -(to a lesser extent) are involved during the paracetamol degradation. Furthermore, heterogeneous catalyst CoO also could react with sulfite to degrade paracetamol to a large extent. In general, this research work provides a new promising strategy by using sulfite and transition metal Co(II) to degrade organic compounds in wastewater under alkaline environment. One of the perspectives of this work is to perform same experiments with other metals in order to assess their ability to promote the radical generation from sulfite under dark and light conditions.

On the basis of the above results and experiments performed in this work, the role of oxygen as oxidant to degrade the aniline directly and accelerating the cycle between Fe(III) and Fe(II) are analyzed accurately. Furthermore, the sequential experiments are also conducted to degrade high concentration of aniline which provide a facile route for decontamination of organic compounds in industrial water with high concentration. Results of this work help to understand the relevance and mechanism of organic contaminants oxidation by SO5 -, which has not been paid much attentions in the conventional SR-AOPs using PMS.

Materials and methods

Chemicals

Ferrous sulfate (FeSO4• 7H2O, analytical reagent grade), ferric sulfate (Fe2(SO4)3), Aniline (AN), Benzotriazole (BTA), N,N-dimethylaniline (DMA), tert-butyl alcohol (TBA), ethanol (EtOH), isopropanol and diphenylamine (DPA), 1,10-phenantroline, NaOH and H2SO4 were purchased from Sinopharm Chemical Reagent Co., Ltd. Sodium sulfite (Na2SO3) from Shanghai Zhanyun Chemical Co., Ltd. was prepared just prior to measurements.

Peroxymonosulfate (2KHSO5. KHSO4. K2SO4 (PMS)), and potassium persulfate (K2S2O8 (PS)) purchased from Sigma-Aldrich, were served as the compounds providing the strong oxidant, peroxymonosulfate ion (HSO5 -). Analytical-grade o-Nitroaniline, p-Nitroaniline, p-aminophenol, p-Chloroaniline, Sulfonamides, Phenol, Paracetamol, NN-dimethylaniline, which were used to testify the possible application of this system were also purchased from Sinopharm Chemical Reagent Co., Ltd. Acetonitrile and methanol were HPLC grade and purchased from Fisher Corporation. Na2CO3 and NaHCO3 were used to determine the concentration of Na2SO3. Ammonium acetate used to prepare buffer solution for Fe(II) titration, were purchased from Sinopharm Chemical Reagent Co., Ltd.

All chemicals were used without further purification. Ultrapure water with 18.2 MΩ cm resistivity used in this work was obtained through a water purification system (Liyuan Electric Instrument Co., Beijing, China).

Experimental setup

Experiments were conducted in a 250 mL open cylindrical reactor held at a constant temperature of 25°C by circulating thermostatically controlled water through an outer water-jacketed. Appropriate amounts of compounds were mixed in the solution, and the pH was adjusted using a pHS-3C pH meter (Hinotek, Ningbo, China) by adding of NaOH or H2SO4 concentrated solutions. Solutions were continuously stirred with a Polytetrafluoroetylene (PTFE)-coated magnetic stirrer, and each run was started by careful addition of the desired volume of a Fe(III) or Fe(II) stock solution. Samples were withdrawn at fixed interval times and analyzed immediately to determine the remaining concentration of Aniline, S(IV) and Fe(II) in solution to avoid further reaction. If necessary, predetermined amounts of radical scavengers were added to the solution before aniline. In order to investigate the effect of oxygen on the aniline degradation, N2 (99.99%) or pure O2 (99.99%)

were bubbled for 30 min before and throughout the experiment using a constant flow rate kept at 1 L min -1 . A dissolved oxygen meter (8403, AZ Instrument Co. Ltd.) was used to determine the concentration of oxygen in solution.

Analytical methods

Aniline disappearance was monitored by high-performance liquid chromatography Shimadzu LC-10A system (Kyoto, Japan) equipped with a UV-vis detector (SPD-10AV; Shimadzu) set at 230 nm, a LC-10AT liquid pump, a SIL-10A auto-injector and a ODS-C18 column (25 cm The concentration of Fe(II) was determined using a colorimetric complexant the 1,10-phenanthroline coupled with spectrophotometric detection at 510 nm using a Shimadzu spectrophotometer UV-1601 [25]. The sulfite concentration was determined by ion chromatography (ICS-90, Dionex, China) using 1mM NaHCO3 and 8 mM Na2CO3 as eluent and dilute H2SO4 as regenerator. The sample loop size was 50 µL and the eluent flow rate was 1.0 mL min -1 .

Results and discussion

Effects of Fe(III)/S(IV) ratio and pH

AN concentration disappearance is monitored in the presence of 0.1 mM of Fe(III) and 1 mM of S(IV) at pH 4.0. No significant degradation is determined when Fe(III) and S(IV) are present separately in solution while, in the presence of both species, AN undergoes significant degradation. In the first 30 minutes about 65 % of AN are degraded while a slowly degradation rate can be observed up to 80 min (Figure 5-1). The initial pseudo-first order degradation rate can be estimated to be (1.32±0.08) × 10 -1 min -1 . In order to investigate the effect of iron and sulfite concentration on degradation, sulfite concentration was modified using 0.1 mM of Fe(III) in solution at pH 4.0. Increasing S(IV) concentration from 1 to 1.2 mM, the AN degradation is strongly inhibited with initial pseudo-first order degradation rate decreasing to (8.16±0.48)×10 -2 min -1 and degradation after 80 min is about 42 %. Increasing the S(IV) concentration inhibits the oxidation of aniline through the scavenging reaction between S(IV) and generated radicals SO4 -or SO5 -.Moreover, decreasing the S(IV) concentration to 0.6 mM no significant change of AN disappearance is noticed (Figure 5-2A). Similar trend was observed keeping a S(IV) concentration constant to 1 mM and varying the Fe(III) from 0.05 to 0.2 mM (Figure 5-2B).

Increasing the Fe(III) concentration may enhance the generation of SO3 -via intermolecular electron transfer, however at high concentration of Fe(III) formation of colloidal Fe(OH)3 particles, which are responsible for inhibition of AN degradation, may occur even at pH 4 [21]. Considering the factors above, we deem 0.1/1.0 (mM/mM) as the optimum Fe(III)/S(IV) molar ratio for this system. Modification of pH strongly affect AN removal using 0.1 mM of Fe(III) and 1 mM of S(IV) as reported in Figure 5-3. At pH 7.0, AN removal is less than 10 % after 80 min while, decreasing the pH the degradation efficiency increased, reaching 70% when the initial pH of solution is 4.0. This effect can be attributed to the stability of Fe(III)-S(IV) complex as reported in different works [14,26]. Although it is rather difficult to summarize the overall influence of pH on Fe(III)-S(IV) induced oxidation (a change in pH can impact the speciation of S(IV) (pKa H2SO3/HSO3 -=1.8) and stability of Fe(III)), it is interesting to note that at pH 4.0 only HSO3are present in solution. Under these conditions high formation of FeSO3 + complexes are expected in agreement with previously reported studies on cobalt reactivity [27].

Determination of the reactivity constants with

BTA SO k M s    (5-2)
Considering a first-order decay of chemical species and applying the competition kinetic method, the relation between the decay of two species gives (Eq 5-1):

4 4 , 00 , [ ] [ ] ln ln [ ] [ ] AN SO tt BTA SO k AN BTA AN k BTA                (Eq 5-1)
Where gives the ratio of two second order rate constants .

Table 1.

Summaries of second-order rate constant k of the reactions of chemical scavengers with oxysulfur radicals.

Reaction

Rate constant, k ( M -1 s -1 )

Reference

EtOH + SO4 •-7.7×10 7 [34,35] EtOH + SO3 •-< 2 × 10 3 [36] EtOH + SO5 •-< 1 × 10 3 [36] EtOH + HO • 1.9 × 10 9 [37] Benzotriazole + SO4 •-

Role of SO4 -and SO5 -

Although the formation of SO4 and SO5 in this system is demonstrated, their relative contribution on the oxidation of AN has been investigated. Under adopted experimental conditions, hydroxyl radical cannot be generated in the system at enough concentration for the oxidation of organic contaminants mainly due to high concentration of S(IV) and acidic environment [24,28]. The fact that no significant inhibition of AN degradation is observed using different TBA concentrations (Figure. [5][6] suggests that HO • radicals are not generated in our system. In order to validate this hypothesis, the initial degradation rate of AN ( d AN R ) is calculated under specific concentration of scavengers, when the concentration of TBA is 100 times higher than those of AN, the corresponding contribution of HO  was less than 3% (the result was shown in In Figure 5-7B the inhibition efficiency is reported as function of ratio [EtOH]0/[AN]0 and the inhibition efficiency can reach ~ 98%, which reasonably shows that complete inhibition of SO4 -is achieved. However, the model fit overestimates the inhibition efficiency. This effect is attributed to the fact that the presence of SO5 -is not considered in the equations.

determine that contribution of SO5 -to aniline degradation is ~60%. From study above, when the concentration of EtOH was 5000[AN]0, the SO4 -could be inhabited exactly without any SO5 -quenched simultaneously. For the sake of accurate differentiating the contributions of various radicals, SO4 -and SO5 -, to AN oxidation, the experiments with variations of the AN initial concentrations (from 10μM to 100μM) were conducted in the presence/ absence of EtOH (when with EtOH added, the concentration was 5000[AN]0) in the solution, and the data was shown in Figure 5-9A and 5-9B. According to the competition kinetics method, the initial rate of Aniline/EtOH with radicals SO4 -and SO5 -can be expressed as Eq 5-6 to 5-7), respectively. Basing on the initial rate fitted at the first reactions stage (0-10min), the carves of inhibited efficiency (ρ) could be calculated out as follows: when the concentration of aniline were 10, 20, 50, 100μM, the inhibited efficiency (ρ) were 38.36%, 46.9%, 42.25%, 37.12% , respectively. Thus, it can be said that the average ρ 41.3% could be the precise concentration of SO4 •in this Fe(III)-sulfite system, this result did not show a big difference to (35.9±1.0) % which obtained in the radical scavengers experiments above. The residual part contribution of 58.7% should belongs to SO5 •-.

Dissolved oxygen implication

To shed light onto the effect of oxygen on aniline oxidation, different experiments are performed at different oxygen concentrations in solution by continuous bubbling pure O2, air and N2 during the whole experiment. As depicted in Figure 5-10, when the experiments are conducted in the presence of air or pure O2 ([O2] > 20 mg L -1 in the solution), a plateau corresponding to 70% of AN degradation is found after 20-30 min. However, the efficiency decreases to 10% in nitrogen purged solutions. This result implies that oxygen plays an important role in transformation of aniline in the Fe(III)-S(IV) system mainly through the generation of SO4 -and SO5 -after reaction with SO3 -in solution. On the other hand, there are not remarkable differences among the degradation effects between bubbling of pure O2 or air after 80 min of reaction, which illustrates that the oxygen bubbling is less effective on PARA degradation only accelerate oxysulfur radical cycle, as shown in reaction 7-10 in Table 5-3. . This trend can be explained considering the fast reactivity with SO3 leading to the formation of SO5 •-(see Table 5-3), which would increase the efficiency of oxidation. In the latter stage of the reaction, sulfite consumption and spontaneous re-oxygenation may explain the slow increase in oxygen content. The pH also changes in the presence of oxygen (Figure 5-11D). H + is mainly generated through reaction between SO4 and HSO3 -(Accordingly to reaction 9-10 in Table 5-3). In saturated oxygen solutions, the reason of higher pH drop was due to the oxysulfur radical cycle acceleration and then enhancement of Fe(III)-S(IV) cycle. On the basis of all the results above, the mechanism of integrated process of this Fe(III)-sulfite system can be shown in Scheme 5-1. To further verify the application of this Fe(III)-S(IV) system, the experiments of degrading various aromatic amines, such as o-Nitroaniline, p-Nitroaniline, Sulfonamides and their substituted analogs within 80 min also were conducted at specific Fe(III)/S(IV) molar ratios (0.1: 1 mM) at initial substrates concentration of 10μM and at pH 4.0. After 80 min reaction, all the pollutants selected in this study underwent different degree of degradation which can be seen from Figure 5-13. As identified above, the application of this system was reasonably proved.

Conclusions

This work shows that the high efficiency in AN decontamination has been achieved in this Fe(III)-S(IV) system under specific conditions. Consequently, oxysulfur radicals including SO4 -and SO5 -are the dominant oxidants which are mainly responsible for aniline degradation. Furthermore, in aniline oxidization, SO5 -also played a relatively important role.

Its contribution takes a considerable part (~ 60%) of all radical species. Additional, the rate constant 𝑘 𝐴𝑁,𝑆𝑂 4 -𝑎𝑛𝑑 𝑘 𝐴𝑁,𝑆𝑂 5 -are estimated to be (7.7 ± 0.5) × 10 9 M -1 s -1 𝑎𝑛𝑑 (5.8 ± 0.6) × 10 6 M -1 s -1 , respectively. When the experiments conducted with different level of oxygen in solution, it shows that dissolved oxygen plays a significant role in radicals generating. This work, not only provides a feasible candidate for use in the efficient oxidation of aromatic amines, but also highlights a novel concept that the role of SO5 -cannot be neglected in the future studies.

Introduction

We already reported the system utilizing Fe(III)-sulfite for degradation of aniline and some other kinds organic compounds in acidic environment in Chapter 5, the main reactions involved in this system are described by reactions 6-1 to 6-10 [1-6]. Although rapid degradation of aniline is achieved, the optimum initial pH for the reaction is about 4.0, and the degradation efficiency is less than 10% at the initial pH 7.0. So An important point of interest is how to enhance the efficiency of the Fe(III)-sulfite system at near-neutral pH. We addressed this problem by exploiting the photochemical activity of ferric iron complexes under UV-vis irradiation [7][8]. The most well-known Fenton system (Fe(II)/H2O2), when exposed to UV-vis is strongly enhanced.

Moreover, the photochemical reduction of ferric species leads to a cycle between Fe(III) and Fe(II) species, and minimizes the precipitation of ferric hydroxide by a large percentage [9][10].

Fe 3+ + HSO3 -↔ FeSO3 + + H + (6-1)

FeSO3 + ↔ Fe 2+ + SO3 - (6-2)
Fe 2+ + HSO3 -→ FeHSO3 + (6-3)

4 FeSO3 + + O2 → 4 FeSO3 + + 2H2O (6-4) SO3 -+ O2 → SO5 - (6-5) SO5 -+ HSO3 -→ SO3 -+ HSO5 - (6-6) SO5 -+ HSO3 -→ SO4 -+ H + + SO4 2- (6-7) SO4 -+ HSO3 -→ SO3 -+ H + + SO4 2- (6-8) SO5 -+ SO5 -→ 2 SO4 -+ O2 (6-9) Fe 2+ + SO5 -→ Fe 3+ + HSO5 - (6-10)
In this study, a detailed discussion on the degradation of paracetamol (PARA), which is chosen to be the organic substrate, in the Fe(III)-sulfite system at near-neutral pH by UV-vis irradiation is reported, and the difference between pH 4.0 and 7.0 with or without UV-vis is also checked to identified and optimized the mechanism for different systems. The effects of initial pH, the Fe(III)/S(IV) molar ratio, oxygen content, buffer are investigated to determine the optimum conditions of the novel photochemical system, and the radical scavengers experiments are conducted to determine the predominant oxidant in such systems. Where Asample and Ablank are the sample and blank adsorption, respectively, ε562 is the molar extinction coefficient, l is the optic path length, Vtot is the total volume and Vsample is the sample volume (1 mL).

Material and methods

Analytical methods

The concentration of PARA is determined using a high-performance liquid chromatography (HPLC) Shimadzu LC-10A system equipped with UV-vis detector (SPD-10AV; Shimadzu) and an ODS-C18 column (25 cm × 4.6 mm, 5 μm; Shimadzu, Kyoto, Japan). The separation is warried out using methanol:water (25:75 v/v) as isocratic mobile phase at a flow rate of 1.0 mL min -1 . The detector is set at 241 nm maximum of absorption of PARA in aqueous solution.

The calibration of Fe(II) is shown in Figure 6 minutes of reaction, the degradation efficiency reached around 69%, and at pH 8.0 or even higher pH, there is still a little extent of PARA degradation. When we compare the efficiency of PARA degradation with or without irradiation, the reaction rate is very fast and could achieve completely the degradation at pH 4.0 with and in a lower extent without light. This result is consistent with earlier reports by our group of greatly enhanced rates of contaminants degradation for Fe(III)-S(IV) system [6,[15][16]. The optimum initial pH of the Fe(III)-S(IV) system is pH 4.0. When the initial pH of the solution increases to 5.0, the initial reaction rate of PARA degradation under irradiation is much faster than without irradiation, but after 60 min of reaction, both of them could reach nearly the same degradation efficiency. This is due to the pH decrease which enhanced the reaction. However, when pH is higher than 6.0, the difference of PARA degradation with or without irradiation become much bigger. For example, at pH = 6.0, the degradation efficiencies are around 39% and 83% without and with irradiation, respectively. This result illustrates that the irradiation greatly enhanced the reaction. The reason can be explained by the pH decreasing to acidic values due to the irradiation. This phenomenon could reveal that photodegradation of organic compounds should be considered as one of the important pathway near-neutral pH environment. respectively. However, from our previous work [6,17], at higher sulfite ions concentrations there is no more enhancement of organic compounds degradation due to the reactivity competition of SO4 -/ SO5 between contaminant and SO3 2-(eqs 6-8 ). In our case, there is no observed inhibition with excess sulfite ions, indicating that photochemical reactions, with the consumption of sulfite, become more significant than the inhibition reactions by excess sulfite.

Considering its potential for causing serious pollution, potential application, and feasibility of use in industrial processes or engineered systems, 1.0 mM is used as the S(IV) dosage when other individual effects are investigated. Batch experiments under irradiation with fixed initial concentration of S(IV) (1.0 mM) are conducted to investigate the effect of initial concentration of Fe(III) in the range of 0.05 to 0.5 mM. Figure 6-6B reveales that the increase of the Fe(III) concentration do not significantly improves the PARA disappearance., On the contrary, when Fe(III) concentration is higher than 0.1 mM, there is an inhibition of PARA degradation. This negative effect is due to the high amount of Fe 2+ generated in the iron cycle and its reaction with sulfate radicals. So these results show that to remove PARA from the aqueous solutions, a small amount of Fe(III) is quite sufficient to obtain good degradation efficiency. So, 0.1 mM of Fe(III) is added to the solution when other individual effects are investigated. One of the reasons explaining the slightly effect of Fe(III) concentration may be the fast recycle of Fe(II) and Fe(III) in the presence of a stoichiometry excess of S(IV) concentrations in aerated solution (eqs 6-1 to 6-2).

Moreover, when the concentration of Fe(III) is monitored during the reaction, it is found that its concentration was nearly stable (Figure 67). Such results suggest that in the presence of enough sulfite in solution, the iron recycle could achieved very fast leading to the generation of SO3 -at the same time.

Effect of oxygen and buffer.

The effect of oxygen on PARA degradation is studied using 0.1 mM Fe(III) and 1.0 mM S(IV), 10 μM PARA at both pH 4.0 and 7.0 with UV irradiation. As shown in Figure 6-8, in the deaerated solution, there is nearly no degradation at pH 7.0 and only 16% loss of PARA at pH 4.0 after 60 min of irradiation is observed. This percentage of PARA degradation is much smaller than the 64% loss at pH 7.0 and complete degradation at pH 4.0 obtained in the aerated solution. This result indicated that oxygen is a very important parameter that affects the photodegradation of PARA in those Fe(III)-S(IV)-O2-hv system at different pH. So oxygen takes part in the photochemical process in such a system. As oxygen plays its role by leading to the generation of SO4 -and SO5 -after reaction with SO3 -in solution, so the phenomenon that there is still some PARA disappearance at pH 4.0 without oxygen can be explained by the fact that there is some hydroxyl radical (HO • ) generated at pH 4.0 under irradiation [18][19][20]. When employed PIPES to be the buffer to keep pH constant at pH 6.7 to 7.0, the result shown in Figure 6-9 indicates that there is nearly no PARA degradation during the whole reaction.

This result is in agreement with the section of pH effect, as the reaction depends heavily on the pH and irradiation can enhanced pH decrease until the reaction happen. However, when using the buffer to keep the pH stable, the system cannot have the appropriate conditions to perform the reaction. To evaluate the reaction mechanism of this system, The radical scavenger experiments are conducted at pH 7.0 with UV, pH 4.0 with UV, and pH 4.0 without UV with fixed concentration of 0.1 mM Fe(III) and 1.0 mM S(IV), 10 μM PARA, using TBA and EtOH to identify the primary effective radicals. As the reaction rate constant of EtOH with HO  is about 1.9 × 10 9 M -1 s -1 , which is about 50-fold greater than with SO4 -, which is around 7.7×10 7 M -1 s -1 and on the other hand, as , TBA HO k  = 6.0 × 10 8 M -1 s -1 [21] which is nearly three orders of magnitude higher than When the experiment is conducted at pH 4.0 with irradiation, the results are shown in Figure 6-10A. 2.0 mM TBA added to the solution results in partial inhibition, and 50 mM EtOH almost completely inhibits PARA degradation. These results illustrate that PARA degradation at pH 4.0 under UV irradiation is due to HO  and SO4 -. The generation of HO  could be divided into two pathways, one of them is due the reaction between SO4 -and H2O (eqs6-11) [22], and another pathway belongs to the photolysis of FeOH 2+ in acidic environment (eqs 6-12 to 6-13) [12]. And the formation of SO4 -come from decomposing FeSO3 + complexes and the oxidation of SO3 -radicals.

SO4 -+ H2O → SO4 For the experiment performs at pH 4.0 in dark and keep other conditions the same, the result shows in Figure 6-10B illustrates that when TBA is added to the reaction solution, no significant inhibition of PARA degradation was observed suggests that HO  are not generated in this system. This fact excludes the hypothesis of HO  formation caused by the reaction between SO4 -and H2O. And when high concentration of EtOH (50 mM) added to the solution, the PARA degradation is nearly inhibited completely. This phenomenon validates that SO4 -is the only species of oxidant that leads to PARA degradation under this condition. For the Fe(III)-S(IV)-O2-hv system at pH 7.0, when EtOH is added to be the scavenger of adducts, respectively described in these references [23][24].In control experiments (Fe(III)-hv system, S(IV) -hv system), there is no signal with the addition 123 of DMPO (results not shown). As a contrary, as shown in Figure 6-12 for the Fe(III) -S(IV)hv system, the hyperfine coupling constants of measured spectra are consistent with those of DMPO/ SO3 -adducts, and no DMPO-OH adducts has been detected. The above phenomenon suggests that HO  is insignificant under this conditions, and SOx -(including SO3 -, SO4 -and SO5 -) is the primary reactive oxidants responsible for PARA degradation.

However, in the absence of DMPO and in the presence of oxygen, SO3 -can be easily oxidized to SO5 -(eq.6-5), thereby considerably mitigating the contribution of SO3 -in our system. Combined with the results of radicals scavenger experiments, EtOH can inhibits completely the reaction, as EtOH has a negligible reactivity with SO5 -(< 1 × 10 3 M -1 s -1 )

[25], it could be concluded that in this Fe(III)-S(IV)-O2-hv system, SO4 -is the oxidants that mainly lead to the degradation of PARA.

Conclusions

The photodegradation of PARA has been investigated in Fe(III)-S(IV)-O2-system in the absence and presence of UV irradiation. The results indicate that pH, Fe(III)/S(IV) concentration and oxygen content all have an important impact on the disappearance of PARA, and to achieve the highest efficiency of PARA degradation, the higher S(IV) concentration should be chosen. Irradiation enhance the PARA degradation via accelerating the pH

2 )

 2 UV/H2O2 processWhen a solution containing H2O2 is irradiated under UV light, the breaking of O-O bond can occur leading to the generation of HO  , eqs. 2-3[10][11][12][13].

Figure. 2 - 2 .

 22 Figure. 2-2. Molecular structure of persulfate

[ 118 -

 118 120]. In early 1900's, accelerated aqueous S(IV) oxidation by O2 in the presence of trace transition metal ion, M n+ , has been observed [121], in the atmospheric liquid phase. Fe(II/III) catalyzes the conversion of S(IV) to S(VI) in the presence of dissolved oxygen. Many investigations of this process have extrapolated detailed mechanisms under various conditions, since it is very important for the generation of acid rain. On the basis of the mechanism of catalysis by iron, much attention has also been paid to flue gas desulfurization, which is an artificially enhanced process of transformation of S(IV) to S(VI) for the purpose of desulfurization. Kinetics and mechanisms have been proposed in eqs 2-35 to 2-37[122-124].

[ 52 ]

 52 Photochemical & Photobiological Sciences Official Journal of the European Photochemistry Association & the European Society for Photobiology, 2011, 10(3): 408-413. [59] Anipsitakis G P, Dionysiou D D. Transition metal/UV-based advanced oxidation technologies for water decontamination [J]. Applied Catalysis B Environmental, 2004, 54(3): 155-163. [60] Jaafarzadeh N, Ghanbari F, Moradi M. Photo-electro-oxidation assisted peroxymonosulfate for decolorization of acid brown 14 from aqueous solution [J]. Korean Journal of Chemical Engineering, 2015, 32(3): 458-464. [61] Olmez-Hanci T, Imren C, Kabdaşlı I, et al. Application of the UV-C photo-assisted peroxymonosulfate oxidation for the mineralization of dimethyl phthalate in aqueous solutions [J]. Photochemical & Photobiological Sciences Official Journal of the European Photochemistry Association & the European Society for Photobiology, 2011, 10(3): 408-413. [62] Herrmann H. On the Photolysis of Simple Anions and Neutral Molecules as Sources of SOx -and Cl in Aqueous Solution [J]. Physical Chemistry Chemical Physics Pccp, 2007, 9(30): 3935-3942.

Figure. 3 - 1 . 3 . 3 BFigure. 3 - 2 (BFigure. 3 - 3 .

 31333233 Figure. 3-1. The calibration of Cr(VI)

Figure 3 -Figure. 3 - 4 .

 334 Figure. 3-4. Species distribution of 0.1 mM Cr(VI) in aqueous solutions at pH in the range of 1 -12.

Figure. 3 - 5 .

 35 Figure. 3-5.Species distribution of Cr(VI) in aqueous solutions at pH 3 containing Cr(VI) at the concentration in the range of 0.1 -2.0 mM. In the concentration range of 0.1 -2.0 mM, HCrO4 -ion is the only predominant species of Cr(VI) (> 90%) at pH 3. The fraction of Cr2O7 2-increases with increasing the total concentration of Cr(VI) and it is about 0.5 when the total concentration of Cr(VI) is at ca. 25 mM. This means in the Cr(VI)-S(IV)-O2 system investigated in this work, HCrO4 -ion is

Figure. 3 - 6 .

 36 Figure. 3-6. Species distribution of 1 mM S(IV) in aqueous solutions at pH in the range of 1 -12.

Figure. 3 - 7 .

 37 Figure. 3-7. Species distribution of S(IV) in aqueous solutions at pH 3 containing S(IV) at the concentration in the range of 0.1 -10.0 mM. In the concentration range of 0.1 -10.0 mM, HSO3ion

[Figure. 3 - 8 .[ 1 [[ 1 BFigure. 3 - 9 .

 381139 Figure. 3-8. Changes in AO7 concentration in the Cr(VI)-S(IV)-O2 system (a) at various initial concentrations of Cr(VI) and (b) at various initial concentrations of S(IV). Conditions: [AO7]0 =

Figure. 3 - 10 .

 310 Figure. 3-10. Change in AO7 concentration in the Cr(VI)-S(IV)-O2 system in the presence of various radical scavengers. Conditions: [AO7]0 = 0.029 mM, [Cr(VI)]0 = 0.1 mM, [Na2SO3]0 = 0.55 mM, pHini = 3.0.

Figure. 3 - 11 . 3 . 3 . 5 Scheme. 3 - 1 .

 31133531 Figure. 3-11. Examples of the regression analysis of the pseudo-first-order kinetic curves of AO7

Figure. 3 - 12 .

 312 Figure.3-12. UV-vis spectra of AO7 at specific time intervals during the reaction. Conditions: [AO7]0 = 0.029 mM, [Cr(VI)]0 = 0.3 mM, [Na2SO3]0 = 1.1 mM, pHini = 3.0.

Figure 3 - 77 Figure. 3 - 13 .

 377313 Figure. 3-13. The pseudo-first-order rate constants((kobs1) versus their associated Hammett σ values which the functional group was substituted at the para and meta position. Substrates: [p-Nitroaniline]0 = [Paracetamol]0 = [Phenol]0 = [p-Chloroaniline]0 = [Bisphenol A]0 = [Aniline]0 = [p-aminophenol]0 = 10μM, [Cr(VI)]0 = 0.1 mM, [Na2SO3]0 = 0.55 mM, pHini = 3.0. The correlation coefficient R is calculated to be 0.77, which is above the critical value of R, 0.755 (N = 7, Significance test (T-test)

Figure. 4 - 1 .

 41 Figure. 4-1. Co(II) calibration curves at 620 nm.

Figure. 4 - 2 .

 42 Figure. 4-2. The effect of control experiments under various conditions on paracetamol degradation. Condition: [PARA]0 = 10μM, [Co(II)]0 = 0.1 mM, [Na2SO3]0 = 1.0 mM, pH0 = 9.0, T =25°C , Fair = 0.8 L min -1 .

from 3 .

 3 Figure.4-3a. Despite a negligible degradation at relatively acidic values (pH 3.0 -5.0), PARA degradation is enhanced from pH 5.0 to pH 10.0 while a decrease is observed at pH 11.0. It is interesting to observe that degradation efficiency between pH 3.0 and 10 is strongly correlated to the HSO3 -/SO3 2-speciation in solution (pKa = 7.2) as shown in Figure.4-3b in which degradation of PARA after 30 min at different pH values is correlated to the concentration of SO3 2-in solution. However, at pH 11, the inhibition of PARA degradation can be attributed to

Figure. 4 - 3 .

 43 Figure. 4-3. (A) Effect of initial pH values on PARA degradation.; (B) Degradation of PARA at different pH values correlated to the concentration of SO3 2-. Condition: [PARA]0 = 10 μM, [Co(II)]0 = 0.1 mM, [Na2SO3]0 = 1.0 mM, T = 25°C , Fair = 0.8 L min -1 .

Figure. 4 - 4 .

 44 Figure.4-4. UV-vis spectra of reaction solution at specific time intervals during the reaction. Conditions: [Co(II)]0 = 0.1 mM, [Na2SO3]0 = 1.0 mM, pH0 = 9.0, T =25°C , Fair = 0.8 L min -1 .

  Different amounts of Co(II) were used to investigate the effect of Co(II) concentration on the PARA removal at pH 9.0 (Figure.[4][5]. When experiments were conducted using a Co(II) concentration between 0.05 and 0.15 mM and a fixed sulfite concentration of 1.0 mM, no significant differences were observed on the PARA degradation profile. One of reasons explaining the slightly effect of Co(II) concentration may be the fast recycle of Co(II) and Co(III) in the presence of a stoichiometry excess of S(IV) concentrations in aerated solution (eqs 4-3 and eqs 4-4). Moreover, when the concentration of Co(II) was monitored during the reaction, it was found that its concentration was nearly stable (Figure.[4][5][6]. Such results

AFigure. 4 - 5 .

 45 Figure.4-5. Effect of Co(II) concentrations on PARA degradation. Condition: [PARA]0 = 10μM, [Na2SO3]0 = 1.0 mM, pH0 = 9.0, T = 25°C , Fair = 0.8 L min -1 .

Figure. 4 - 6 .Figure. 4 - 7 .

 4647 Figure.4-6. Co(II) concentration during the reaction. Condition: [PARA]0 = 10μM, [Co(II)]0 = 0.1 mM, [Na2SO3]0 = 1.0 mM, pH0 = 9.0, T =25°C , Fair = 0.8 L min -1 .

( 5 .Figure. 4 - 8 .

 548 Figure. 4-8. Effect of different initial concentration of PARA. Inset: the initial reaction rate for different paracetamol concentration. Condition: [Co(II)]0 = 0.1 mM, [Na2SO3]0 = 1.0 mM, pHini= 9.0, T =25℃,Lair= 0.8L/min.

Figure. 4 - 9 .

 49 Figure. 4-9. Effect of oxygen on PARA degradation. Condition: [PARA]0 = 10μM, [Co(II)]0 = 0.1 mM, [Na2SO3]0 = 1.0 mM, pH0 = 9.0, T = 25°C , Fair = 0.8 L min -1 .

Figure. 4 - 10 .

 410 Figure. 4-10. Pseudo-first order decay of sulfate radical monitored at 470 nm vs PARA concentration

Figure. 4 - 11 .Figure. 4 - 12 .

 411412 Figure.4-11. Effect of TBA on paracetamol degradation. Condition: [PARA]0 = 10μM, [Co(II)]0 = 0.1 mM, [Na2SO3]0 = 1.0 mM, pH0 = 9.0, T =25°C , Fair = 0.8 L min -1 .

Figure. 4 -

 4 Figure. 4-13 and Figure. 4-5, the initial degradation rate is completely different in the two systems, very fast with soluble Co(II) and much slower with CoO. This observation showsthat the reactivity of sulfite with cobalt is efficient in homogeneous phase and so with cobalt oxide a first process of solubilisation seems necessary.

Figure. 4 - 13 . 1 4. 3 . 8 Figure. 4 - 14 .

 413138414 Figure. 4-13. Effect of heterogeneous catalyst CoO react with S(IV) to degrade PARA. Condition: [PARA]0 = 10μM, [Na2SO3]0 = 1.0 mM, pH0 = 9.0, T = 25°C, Fair = 0.8 L min -1

× 4 .

 4 6 mm, 5μm; Shimadzu, Kyoto, Japan). The eluent was mixture of acetonitrile/water (v/v, 55/45) with a flow rate of 1.0 mL min -1 . The injection volume was 20 μL and the retention time for aniline was 4.7 min in these experimental conditions. The condition of monitoring the residual of other aromatic amines by HPLC were as follow: o-Nitroaniline: 230nm with water/acetonitrile mixture (45/55 v/v); p-Nitroaniline: 272nm with water/methanol mixture (45/55 v/v), p-Chloroaniline: 260nm with water/acetonitrile mixture (45/55 v/v), Sulfonamides: 270nm with water/acetonitrile mixture (45/55 v/v), Phenol: 270nm with water/methanol mixture (45/55 v/v), Paracetamol: 241nm with water/ methanol mixture (75/25 v/v), p-aminophenol: 225nm with water/methanol mixture (50/50 v/v), NN-dimethylaniline: 254 nm with water/acetonitrile mixture (45/55 v/v).

3 Figure. 5 - 1 .

 351 Figure. 5-1. Degradation of aniline in the control experiments under various conditions under aerated solution. Condition: [AN]0 = 10μM, [Fe(III)]0 = 0.1mM, [S(IV)]0 = 1.0 mM, pHini = 4.0.

BFigure. 5 - 2 .

 52 Figure. 5-2. Degradation of AN (10 µM) using different Fe(III) /S(IV) ratios at pH = 4.0 under aerated solution. Condition: [AN]0 = 10μM, pHini = 4.0.

Figure. 5 - 3 .

 53 Figure. 5-3. Degradation of AN using 0.1 mM of Fe(III) and 1 mM of S(IV) at different pH values under aerated solution. Condition: [AN]0 = 10μM, [Fe(III)]0 = 0.1mM, [S(IV)]0 = 1.0 mM.

4 SO  and 5 SO

 45 The second-order rate constants between AN and SO4 -chemical competition kinetic with radical scavengers as listed in Table5-1. Bimolecular reactivity constant with SO4 -is determined in a conventional Fe(II)-PS system in which SO4 -and OH  generated at initial pH 3. Isopropanol is used to scavenge HO  while Benzotriazole (BTA) is employed as chemical competitor for the reaction with SO4 -(eqs 5-1 and eqs 5-

Figure. 5 - 4 .Figure. 5 - 5 .

 5455 Figure. 5-4. Linear fit of ln([AN]0/[AN]t) vs ln([BTA]0/[BTA]t) in the system Fe(II)-S2O8 2in the

Figure. 5 - 6 . 2 ( 1 -

 5621 Figure. 5-6. Inhibition of the oxidation of aniline by TBA in Fe(III)-S(IV) system. Conditions: [aniline]0 = 10μM, [Fe(III)]0 = 0.1mM, [S(IV)]0 = 1.0 mM, pHini = 4.0.Table 2 Initial degradation rate of AN ( d AN R ) in the Fe(III)-S(IV) system in the presence of various

  of [EtOH] (ρ) on Aniline oxidation by SO4 -can be determined as:

Figure. 5 - 8 .

 58 Figure. 5-8. AN degradation in the system Fe(III)-S(IV) in absence (no scavengers) and presence of different radical scavengers. Conditions are [AN]0 = 10 μM, [Fe(III)]0 = 0.1 mM, [S(IV)]0 = 1.0 mM, pH = 4.0.

𝑟 0 ,

 0 𝐴𝑁,𝑆𝑂 5 -= 𝑘 𝐴𝑁,𝑆𝑂 5 -[𝐴𝑁] 0 [𝑆𝑂 5 -] 𝑠𝑠 (Eq.5-6) 𝑟 0,𝐸𝑡𝑂𝐻,𝑆𝑂 4 -= 𝑘 𝐸𝑡𝑂𝐻,𝑆𝑂 4 -[𝐸𝑡𝑂𝐻] 0 [𝑆𝑂 4 -] 𝑠𝑠 (Eq.5-7) 𝑟 0,𝐸𝑡𝑂𝐻,𝑆𝑂 5 -= 𝑘 𝐸𝑡𝑂𝐻,𝑆𝑂 5 -[𝐸𝑡𝑂𝐻] 0 [𝑆𝑂 5 -] 𝑠𝑠 (Eq.5-8) Furthermore, we define inhibited efficiency (ρ) of initial rate of aniline by excessive EtOH as: r0,EtOH and r0 represent the initial rate of aniline in the presence and absence of excessive EtOH, respectively.

Figure. 5 - 9 .

 59 Figure. 5-9. The effect of initial aniline concentration during oxidation in Fe(III)-S(IV) system

Figure. 5 - 10 .

 510 Figure. 5-10. Concentration changes of aniline (C/C0) during oxidation in Fe(III)-S(IV) system with or without oxygen. Conditions are [AN]0 = 10 μM, [Fe(III)]0 = 0.1 mM, [S(IV)]0 = 1.0 mM, pH = 4.0.

Figure. 5 - 11 .

 511 Figure. 5-11. Variation of S(IV) (A), Fe(II) (B), [O2] (C), and pH (D) during oxidation of AN in Fe(III)-S(IV) system. Conditions are: [AN]0 = 10 μM, [Fe(III)]0 = 0.1 mM, [S(IV)]0 = 1.0 mM, pH =

Figure. 5 -

 5 Figure. 5-11C shows that oxygen concentration sharply declined in the first 10 min before to

Figure. 5 - 12 .Figure. 5 - 13 .

 512513 Figure. 5-12. AN 100 µM degradation during oxidation in Fe(III)-S(IV) with sequential addition of

Figure. 6 - 1 .FeFe

 61 Figure. 6-1. Emission spectrum of Xenon lampFe(II) and Fe(III) concentrations are determined by complexation with Ferrozine[11]. Firstly, Fe(II) is determined measuring the absorption at 562 nm of the solution obtained by mixing 1 mL of sample, 4 mL of water and 500μL of Ferrozine 20mM. As shown in eqs 6-1, the molar absorption coefficient of the complex at 562 nm is 27900 M -1 cm -1 .Fe(II) =𝐴𝑠𝑎𝑚𝑝𝑙𝑒-𝐴𝑏𝑙𝑎𝑛𝑘 𝜀562 ×𝑙

Figure. 6 - 2 .

 62 Figure. 6-2. Calibration curves of (A) Fe(II); (B) Fe(III) at 562 nm

4 AFigure. 6 - 9 BFigure. 6 -

 4696 Figure. 6-5A. Effect of pH on PARA degradation without UV irradiation under aerated solution . Condition: [PARA]0 = 10 μM, [Fe(III)]0 = 0.1 mM, [S(IV)]0 = 1.0 mM.

Figure. 6 -Figure. 6 -

 66 Figure. 6-6A. Effect of S(IV) concentration on PARA degradation with UV irradiation under aerated solution. Condition: [PARA]0 = 10μM, [Fe(III)]0 = 0.1 mM, pHinit = 7.0.

Figure. 6 - 7 .

 67 Figure. 6-7. Fe(III) concentration during the reaction under aerated solution. Condition: [PARA]0 = 10 μM, [S(IV)]0 = 1.0 mM, [Fe(III)]0 = 0.1 mM, pHinit = 7.0.

7 Figure. 6 - 8 .

 768 Figure. 6-8. Effect of oxygen on PARA degradation with UV irradiation. Condition: PARA]0 = 10 μM, [S(IV)]0 = 1.0 mM, [Fe(III)]0 = 0.1 mM.

Figure. 6 - 9 .6. 3 . 5

 6935 Figure. 6-9. Effect of buffer on PARA degradation with UV irradiation under aerated solution. Condition: [PARA]0 = 10 μM, [S(IV)]0 = 1.0 mM, [Fe(III)]0 = 0.1 mM, pHinit = 7.0 6.3.5 Effect of radical scavengers.

 = 8 . 5 ×

 85 10 5 M -1 , hence on the basis of these properties, by using TBA and EtOH, different inhibitory effects can be observed on the PARA degradation.

Figure. 6 -

 6 Figure. 6-10A. Effect of radical scavengers on PARA degradation with UV irradiation under aerated solution. Condition: PARA]0 = 10 μM, [S(IV)]0 = 1.0 mM, [Fe(III)]0 = 0.1 mM, pHini = 4.0.

Figure. 6 -

 6 Figure. 6-10B. Effect of radical scavengers on PARA degradation without UV irradiation under aerated solution. Condition: [PARA]0 = 10 μM, [S(IV)]0 = 1.0 mM, [Fe(III)]0 = 0.1 mM, pHinit = 4.0.

SO4Figure. 6 -Figure. 6 -

 66 Figure. 6-11A. Effect of EtOH on PARA degradation with UV irradiation under aerated solution. Condition: [PARA]0 = 10 μM, [S(IV)]0 = 1.0 mM, [Fe(III)]0 = 0.1 mM, pHinit = 7.0. 0

Figure. 6 - 12 .BFigure. 6 - 14 .BFigure. 6 - 15 .

 612614615 Figure. 6-12. ESR spectrum of reaction solution. Condition: [S(IV)]0 = 1.0 mM, [Fe(III)]0 = 0.1 mM.[DMPO] = 1g L -1 .

Table 3 -1.

 3

	Pseudo-first-order Rate (kobs) of AO7 in the Cr(VI)-S(IV)-O2 System in the Presence of
	Various Radical Scavengers
	parameter	none	concentration ratio of radical scavengers [TBA]0/[AO7]0 = 400 [aniline]0/[AO7]0 = 100
	kobs (min -1 )	0.120	0.094	0.007
	contribution of HO • + SOx •-		(1 -kobs,aniline,AO7/kobs) × 100 = 94.2%
	contribution of HO •		(1 -kobs,TBA,AO7/kobs) × 100 = 21.7%
	contribution of SOx •-a		93.4% -22.8% = 72.5%
	contribution of other oxidant		100% -94.2% = 5.8%

Table 3 - 2

 32 Degradation efficiency of various kinds of organic pollutants used in this work

	Compounds	Degradation efficiency at 30	Degradation efficiency at 60
		min		min
	Aniline	38.2 ± 1.7 %	42.2 ± 1.2 %
	o-Nitroaniline	24.1 ± 0.2 %	25.0 ± 0.5 %
	p-Nitroaniline	13.3 ± 0.5 %	13.8 ± 0.7 %
	p-Chloroaniline	37.4 ± 0.8 %	38.0 ± 0.6 %
	Sulfonamides	42.6 ± 0.7 %	43.1 ± 0.1 %
	Phenol	64.9 ± 0.6 %	65.9 ± 1.7 %
	Bisphenol A	54.9 ± 1.1 %	56.7 ± 0.1 %
	2-Bromophenol	71.5 ± 1.9 %	75.3 ± 1.6 %
	Paracetamol	46.2 ± 1.0 %	46.3 ± 1.4 %
	p-aminophenol	80.8 ± 1.4 %	84.1 ± 1.0 %
	Catechol	73.5 ± 0.2 %	75.4 ± 0.3 %
	o-Nitrophenol	58.7 ± 1.8 %	61.5 ± 2.4 %
	2,6-Dinitrophenol	4.80 ± 0.2 %	5.90	± 0.1 %

Table 3

 3

	-2, the reactions after 30 min significantly

Table 3 - 3 .

 33 The pseudo-first-order rate (kobs1) and the Hammett σ values for the compounds

	Compounds	kobs1( min -1 )	σ
	p-Nitroaniline	0.06737	0.58
	o-Nitroaniline	0.04746	0.11
	p-Chloroaniline	0.03277	-0.45
	Aniline	0.0321	-0.67
	Sulfonamides	0.02844	-0.07
	2,6-Dinitrophenol	0.10129	1.18
	o-Nitrophenol	0.01775	0.4
	2-Bromophenol	0.01119	-0.15
	Paracetamol	0.02578	-0.38
	Phenol	0.01442	-0.38
	Bisphenol A	0.01998	-0.58
	Catechol	0.01026	-0.75
	p-aminophenol	0.00712	-1.04

Table 5

 5

				-2). To dissociate the contribution of SO4 -	and SO5 -	, different
	scavengers are added into the reaction system. Scavenging of SO3 -	by EtOH even at very
	high concentration (0.5 M) can be neglected due to the very high reactivity between SO3 -	and
	dissolved oxygen (	k	 = 2.3×10 9	M -1	s -1	[29]). However, EtOH can be used to differentiate
			23 , O SO
	SO4 -	and SO5 -	due to large difference (more than 10 4	-fold) between
							, EtOH SO
							4

  2-+ H + + HO •(6)(7)(8)(9)(10)(11)

	FeOH 2+ + UV → Fe 2+ + HO 	(6-12)
	FeSO3 + + UV → Fe 2+ + SO3 -	(6-13)

Acknowledgements I would like to thank gratefully to my supervisors Prof. Marcello BRIGANTE, Prof. Gilles. MAILHOT and Prof. Feng WU for giving me the chance to conduct the thesis work. It is my pleasure to work with all of them and I learned a lot of thing from them. I would also like to thank Engineer Guillaume VOYARD and Prof. Jinjun LI for giving me precious advice and technical supports. I wish to express my gratitude to Prof. Khalil HANNA (Ecole Nationale Supé rieure de Chimie de Rennes, France), Zongping WANG (University of Science and Technology, Chine) and Corinne FERRONATO (Université Claude Bernard Lyon 1, France) for reading and evaluating my thesis. I also wish to thank Prof. Hui ZHANG (Wuhan University) and Prof. Jinjun LI (Wuhan University) for their acceptation to be part of thesis jury. I also wish to thank all my colleagues in Advanced Oxidation Lab of Wuhan University and Institut de Chimie de Clermont-Ferrand of Université Clermont Auvergne for their kindness help. Thank for the fund support of China Scholarship Council (CSC) affiliated with the Misistry of Education of P.R. China.

Finally, I would especially like to thank my parents and my husband for always been there with encouragement.

Sulfate radical (SR, SO4 -) based advanced oxidation process (AOPs), namely SR-AOPs, has currently drawn much attention in the field of oxidative decontamination of polluted waters and soils [1][2][3][4][5]. Persulfate (PS, S2O8 2-) and peroxymonosulfate (PMS, HSO5 -) have been used as SO4 -precursors in previously reported investigations [6][7][8][9][10][11][12][13][14] using homogeneous or heterogeneous catalysts containing transition metal such as Fe(II) and Co(II) [15,16].

Recently, bisulfite (HSO3 -) has been used to replace PS or PMS for new transition-metal-catalyzed SR-AOPs systems for oxidation of several contaminants such as azo dyes [14,[17][18][19], phenols, aromatic amines, and arsenite [20][21][22]. Mechanism of radical chain reactions involving oxysulfur radicals (mainly sulfite radical (SO3 -), peroxymonosulfate radical (SO5 -), and SO4 -) during S(IV) autoxidation have been investigated. Generally, SO5 -is formed from the reaction of SO3 -with molecular oxygen leading to the formation SO4 -through further reaction with another SO5 -. However, the involvement of SO5 -during the oxidation of contaminants has not been completely demonstrated. Neta and Huie reported one-electron redox reactions between SO5 •-and aromatic amines, hydroquinone and other hydroxyphenols [23]. The authors suggested that SO5 -and not SO3 -was responsible for the observed degradation of aniline in solution. On the basis of the oxidation potential, SO5 -can be used to achieve the degradation of organic compounds. Moreover, SO5 -can improve the Fe(III)/Fe(II) cycle, even if the mechanisms remain unclear. In the absence of oxygen, nearly no oxidation of organic contaminants has been observed, which exclude the role played by SO3 - [20,21]. The evidence of Fe(III)/Fe(II) cycle in the presence of saturated oxygen solutions was provided by Brandt et al., [24]. The authors suggested that SO5 -could be the main oxidant during Fe(II) oxidation into Fe(III).

Since the formation of SO5 -is the main oxygen-consuming step during the overall redox process in Fe(III)-S(IV) system, the role of oxygen during the regeneration of Fe(III) from Fe(II) is not still elucidated as well as the enhancement of the contaminant oxidation through the activation by iron of S(IV).

In this work, the mechanism of the aniline (AN) (used as contaminant compound model) oxidation is investigated in the Fe(III)-S(IV) system, with emphasizing the dual role of oxygen. The changes in concentrations of Fe(III) and sulfite, initial pH value of reaction solution are investigated; the second-order rate constants between aniline and SO4 -/SO5 -are also measured, then the contribution made by SO4 -and SO5 -are all evaluated in this work. DPA and EtOH are also used in order to discriminate the contribution of SO4 -and SO5 -to the AN degradation in the Fe(III)-S(IV) system. In the presence of 50 µM of DPA that is able to quantitatively trap both SO4 -and SO5 -, almost no degradation of AN is observed in solution as reported in Figure . 5-8 indicating that SO4 -and SO5 -are the most important species generated in solution. Using 50 mM of EtOH which is able to scavenge SO4 -and not significantly SO5 -, AN degradation is inhibited of about 50 %.

The contribution of these two radicals can be then determined and, as shown in Table 2, the inhibition efficiency accounted by for SO4 -is only (35.9±1.0) % and it is possible to

Referring to the consumption of S(IV) which is shown in Figure 5-11A sulfite ions are oxidized to S(VI) very quickly especially in oxygen satured solution. The reason for decline of S(IV) amount under anoxic conditions is due to the formation of SO3 -(equilibria 1 and 2 in Table 5-3). Under oxic conditions, rapid increase in S(IV) oxidation within several minutes may be explained by rapid formation of FeSO3 + [14,30]. In figure 5 (corresponding to 7 times 1 mM) were added to the solution in one step, the reaction is strongly inhibited. In fact, an increase in S(IV) dosage is not always able to enhance organic contaminants oxidation. On the contrary, excessively high S(IV) concentrations retards the subsequent reaction as it is observed in Figure 2A and in different references [17,31]. This results is in agreement with previously reported works on high [M(n)]/[S(IV)] and low [S(IV)]/[O2] ratios (M(n): heavy-metal ion) [32,33]. Otherwise, the system tends to predominantly undergo reduction. However, as the source of generating oxysulfur radicals, an appropriate concentration of S(IV) can increase the oxidation reaction.

Chapter 6 Enhanced degradation of paracetamol by the Fe(III)-sulfite system under UV irradiation 6.3 Results and discussion

Control experiments

To evaluate the performance of the Fe(III)-S(IV)-O2-hv system, PARA is selected to be the model pollutant, and its degradation in aqueous solution at near-neutral pH is investigated. This phenomenon suggests that Fe(III)-S(IV) complex, FeSO3 + could be formed first [6,7

12-13], then the FeSO3 + complex might have undergone LMCT (ligand-to-metal charge transfer) [14] and SO3 -could be generated as described in Zhang et al. report [7]. On the other hand, when the change of pH is checked during the reaction, the result shown in Figure 6-4 indicates that when there is irradiation, pH will decrease to a great extent during the reaction. At the first stage of the reaction, pH decreases slightly, then after 10-20 min reaction, the pH decreases sharply. When comparing the trend of PARA degradation, it is found that the speed of degradation of PARA has the same trend with the pH decreasing. From the result above, it can be seen that irradiation can enhance the reaction at near-neutral pH in this Fe(III)-S(IV)-O2-hv system due to the stronger decrease of pH until an acidic environment (pH = 3.5).

Effect of pH on PARA degradation

To determine the effect of pH on PARA degradation, samples are analyzed at initial pH values of 4.0 to 9.0 with or without irradiation. Figure 6-5A shows that when there is no irradiation the efficiency of PARA degradation decreases with the increase of initial pH values. When the initial pH is at 6.0, the degradation efficiency of PARA is near just 39% after 60 min of reaction. And when the pH value of the solution is equal or higher to 7.0, there is nearly no PARA degradation. Corresponding experiments with irradiation are analyzed to verify the effect of illumination on PARA degradation in the same pH range between 4.0 to 9.0. As shown in Figure 6-5B, under irradiation even the efficiency of PARA degradation still decreases with the increase in pH values. However, the reaction can happen at pH 7.0, after 60

Scheme. 6-1. The proposed mechanism for Fe(III)-S(IV) system at different pH values.

Effect of transition metal ions-S(IV)-O2-system on PARA degradation

In order to investigate whether the other transition metal ions also has the same function as Fe(III), Cr(VI), Co(II), Cu(II) are chosen to react with sulfite under UV irradiation at near-neutral pH. From previous reports [22], Cr(VI) can work with S(IV) in dark under acidic environment, and pH 3.0 -3.5 is the optimum condition for the reaction. In this case, using 365 nm UV to be the light source, the results in Figure 6-13 show that Cr(VI)-S(IV)-O2system do not have ability to degrade PARA at initial pH 7.0 even with UV irradiation. The concentration of PARA start to decrease after 40 min reaction. Combined with the change of solution pH, the trend is totally the same than with Fe(III)-S(IV)-O2-hv system, expect for the speed of decreasing pH that is much lower. From Figure 6-13, we can conclude that when the pH value of the reaction solution is decreased lower than 5.0, the reaction can start. In the case of the experiments are conducted without UV irradiation, after 60min reaction the pH decreases at a value lower than 5.0 and then PARA starts to degrade. From the result above, it can be concluded that this Cr(VI)-S(IV)-O2-hv system also strongly depends on the pH, and irradiation enhances the PARA degradation via enhancing pH decrease. degradation by Fe(III)-S(IV)-O2-system with initial pH around 7.0, these attributes make the Fe(III)-S(IV)-O2-hv system a more powerful system to deal with the wastewater at near-neutral pH.

Chapter 7 General conclusion and future perspective

General conclusion

In this thesis, sulfite activatio by transition metal ions Cr(VI), Co(II), Fe(III) to generate oxidant species such as oxysulfur radicals (SO3 -, SO4 -and SO5 -) and hydroxyl radical (HO  ) was investigated. The results obtained during this thesis work concern the optimum conditions, the removal efficiency and the chemical mechanism investigation of all the systems.

(1) The Cr(VI)-S(IV)-O2 system is an efficient system for rapid and simultaneous reduction of Cr(VI) and oxidative degradation of organic pollutants in aqueous solution with dioxygen under acidic conditions. We have considered effects of pH, concentration of added Cr(VI) and S(IV), as well as the presence of O2 or N2 and various radical scavengers. Our results show that the reaction of Cr(VI) and degradation of organic pollutants depended heavily on the pH and initial concentrations of Cr(VI) and S(IV). In addition, oxysulfur radicals were the dominant oxidants for AO7 degradation. Thus, our Cr(VI)-S(IV)-O2 system may be a new candidate for use in the efficient treatment of waste using waste chromium via simultaneous reduction of Cr(VI) and oxidation of organic contaminants.

(2) High efficiency of the Co(II)-S(IV)-O2 system using PARA as organic pollutant model in aqueous solution at alkaline pH was achieved. In this system, the degradation of paracetamol depend on initial concentrations of S(IV) and strongly the pH. In fact, it is clear that dissolved oxygen plays a crucial role allowing the oxysulfur radical oxidation to initiate the reaction. The results of radical scavenger experiments demonstrate that SO4  -and also SO5  -(to a lesser extent) are involved during the paracetamol degradation.

Furthermore, heterogeneous catalyst CoO also could react with sulfite to degrade paracetamol with high efficiency. In general, this research work provides a new promising strategy by using sulfite and transition metal Co(II) to degrade organic compounds in wastewater under alkaline environment. One of the perspectives of this work is to perform same experiments with other metals in order to assess their ability to promote the radical generation from sulfite under dark and light conditions.

(3) High efficiency in AN decontamination has been achieved in this Fe(III)-S(IV)-O2 system under specific conditions. Consequently, oxysulfur radicals including SO4 -and SO5 -are the dominant oxidants which are responsible for aniline degradation. Furthermore, in aniline oxidation, SO5 -also played a relatively important role, the contribution of which took the considerable part (~ 60%) of all species of radicals. Additional, the rate constant 𝑘 𝐴𝑁,𝑆𝑂 4 -𝑎𝑛𝑑 𝑘 𝐴𝑁,𝑆𝑂 5 -were estimated to be (7.7 ± 0.5) × 10 9 M -1 s -1 𝑎𝑛𝑑 (5.8 ± 0.6) ×

Perspectives

Using S(IV) to be the reductant, O2 to be the oxidant and transition metal ions to be the initiator, those transition metals-sulfite processes can open a new perspective of oxidative reactions to degrade the organic pollutants. Discovering novel methods for activating S(IV) have been recently proliferated indicating these transition metals + sulfite processes as new tools for remediation of contaminated environments. However, it should be stated that those processes are neoteric processes which still have a lot to go through. Future studies should be developed as follows:

(1) Finding out more kind of metal ions that can react with sulfite to form oxysulfur radicals, and find out the similarity for all those metal ions, the understanding of behavior of the metals is necessity for which further efforts and studies are required.

(2) As TOC removal efficiency were very low in all those systems, so try to find out more method to make the metal ions/S(IV) systems works in alkaline environment, as it facilitate the generation of HO • which have the capacity to mineralize organic compounds.

Furthermore, coupling of this transition metals + sulfite system with other AOPs for synergistic effect may also be an excellent strategy in the efficient degradation of organic compounds and achieve desired TOC removal efficiency.

(3) All the solutions contained organic pollutants are synthetic wastewater which are prepared in the lab, in the future study, we can use those systems to degrade target pollutants in actual wastewater. Actually, their application in industrial scale requires many studies to stabilize their position. Howbeit, due to desirable performance of this transition metals + sulfite systems in degrading the pollutants, it is not an exaggeration to apply those processes in industrial scale in the near future.