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Self-enhanced ozonation of benzoic acid at acidic pHs
w a t e r r e s e a r c h 7 3 ( 2 0 1 5 ) 9 e1 6
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Self-enhanced ozonation of benzoic acid at acidic pHs Xianfeng Huang a, Xuchun Li a,b, Bingcai Pan a,*, Hongchao Li a, Yanyang Zhang a, Bihuang Xie a a
State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing 210023, China b School of Environmental Science and Engineering, Zhejiang Gongshang University, Hangzhou 310012, China
article info
abstract
Article history:
Ozonation of recalcitrant contaminants under acidic conditions is inefficient due to the
Received 2 September 2014
lack of initiator (e.g., OH) for ozone to produce hydroxyl radicals (HO). In this study, we
Received in revised form
reported that benzoic acid (BA), which is inert to ozone attack, underwent efficient
2 January 2015
degradation by ozone at acidic pH (2.3). The kinetics of BA degradation and ozone
Accepted 7 January 2015
decomposition were both enhanced by increasing BA concentrations. Essentially, it is a
Available online 16 January 2015
HO-mediated reaction. Based on the exclusion of possible contributions of H2O2 and
Keywords:
mation of ozone ion (O3 $ ), which is an effective precursor of HO, was thus proposed. The
Ozone
hydroxycyclohexadienyl-type radicals generated during the attack of BA by HO may lead
Acidic conditions
to the formation of O3 $ . Meanwhile, O3 $ could also be possibly formed from the reaction
Benzoic acid
between ozone and organic (e.g., ROO∙) or inorganic peroxyl radicals (e.g., HO2). In addi-
phenol-like intermediates for HO production, the reaction mechanism involved the for-
tion, the hydroxylated products like phenol-like intermediates also played a positive role in
HO
Advanced oxidation process
HO production. Consequently, HO was produced efficiently under acidic conditions, resulting in rapid degradation of BA. This study provides a new approach for ozone activation even at acidic pHs, and broadens the knowledge of ozonation in removal of micropollutants from water. © 2015 Elsevier Ltd. All rights reserved.
1.
Introduction
Nowadays, an increasing discharge of acidic wastewater from various industries, e.g., pesticide, printing and dyeing, pharmacy, chemical industry and electroplating, shows potential risks to ecosystem and human health (Gogate et al., 2004; Oller et al., 2011). A variety of organic compounds therein, including pesticides, pharmaceuticals, PCBs, PAHs, nitroaromatics, and dyes, need advanced treatment due to their carcinogenic, * Corresponding author. Tel.: þ86 25 8968 0390. E-mail address: [email protected] (B. Pan). http://dx.doi.org/10.1016/j.watres.2015.01.010 0043-1354/© 2015 Elsevier Ltd. All rights reserved.
teratogenic and mutagenic effects (Ikehata et al., 2008; Sirtori et al., 2009; Kim et al., 2011). However, efficient and safe removal of these contaminants from acidic wastewater is still challenging now. Currently, several methods have been developed and widely used to treat acidic wastewater, such as biological process, membrane separation, and advanced oxidation (AOPs) (Oller et al., 2011; Marcucci et al., 2001; Gogate et al., 2004). The hydroxyl radical (HO)-involved AOPs, including
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Fenton oxidation, wet oxidation, photochemical oxidation and electrochemical oxidation, are very promising and have been attracting great interests. Despite their different advantages in water treatment, some inherent disadvantages have to face now. In brief, Fenton oxidation has to face the high reagent dosage, iron sludge yield and hydrogen peroxide residue (Pham et al., 2009). Wet oxidation is usually performed under high temperature and pressure. Thus, the reaction equipment should be of resistance to high temperature/pressure and corrosion, resulting in high investment and operation cost (Kim and Ihm, 2011). In addition, most of the photochemical oxidation is still far from real application because it is inevitably restricted to short light penetration distance in wastewater and high energy cost (Yang et al., 2014). Also, electrochemical oxidation still encounters some problems such as the electrode corrosion and passivation and low current efficiency (Panizza et al., 2000). In general, further improvement is still required to forward the above AOPs to wider application in wastewater treatment. Comparatively, ozonation is widely used for taste and odor control, decoloration, elimination of organic and inorganic contaminants, by direct oxidation using molecular ozone and/or through radical reactions involving HO (Von Gunten, 2003). The HO mediated ozonation is attractive and promising in removal of recalcitrant contaminants, such as nitrobenzene, pesticides, pharmaceuticals, and even persistent organic pollutants (Ikehata et al., 2008; Buffle et al., 2006). However, the production of HO from ozone needs initiator like hydroxide ions (OH). Thus, its efficiency is dramatically decreased under acidic conditions. To improve the efficiency of HO generation during ozonation at acidic pHs, some effective strategies have been proposed, including combination of H2O2, energy input (e.g., UV and ultrasound), and addition of transition metal ions or transition metal oxides (Kasprzyk-Horden et al., 2003). Recently, some electron-rich moieties (ERMs) such as phenols, amines, alkoxybenzenes, and hydroxylation product formed intermediately during ozonation have been identified to be capable of activating O3 to produce HO radicals (Buffle et al., 2006b; Mvula et al., 2009). On the contrary, little attention was paid to O3 activation by electron-deficient moieties (EDMs) such as aromatics compounds substituted by electron-withdrawing groups (EWG) or their intermediate species. This is possibly because they are usually ozonerefractory. On the other hand, compounds substituted with EWG such as eCOOH, eNO2, eCl, and eCN could be efficiently oxidized by HO radicals at second-order rate constants from ~108 to ~1010 M1 s1, accompanying the formation of hydroxycyclohexadienyl-type radicals (HO-adduct radicals) (Merga et al., 1996; von Sonntag, 2006). Those HO-adduct radicals can serve as oxygen-transfer media or electron shuttle (Von Sonntag, 2006; Pan et al., 1993; Duesterberg et al., 2007), and thus react with a variety of oxidants such as O2, quinone, and metal ions (e.g., Fe3þ, Cu2þ, Cr6þ) (Steenken et al., 1979; Raghavan et al., 1980). The HO-adduct radicals also react irreversibly with O3 via oxygen transfer based on quantum-chemical calculation (Naumov and von Sonntag, 2011). Considering much stronger oxidizing ability of O3
than the above oxidants, the HO-adduct radicals may participate in the formation of ozone ion (O3 $ ), a precursor of HO, and the enhanced production of HO by those radicals in ozonation under acidic conditions could be anticipated. However, the efficiency of ozone activation by aromatics substituted with EWG or their intermediates at acidic pHs is still unclear. The primary objective of this study was to investigate the efficiency of ozone activation by aromatics substituted with EWG or their intermediates at acidic pHs, benzoic acid being employed as a representative compound. Kinetics of BA degradation and ozone decomposition at acidic pH (2.3) were investigated. Effects of the concentration of BA and ozone dosage on the HO production efficiency were evaluated. Electron paramagnetic resonance studies (EPR) was used to identify the production of HO. By analyzing the roles of reactive species intermediately generated in the activation of ozone as well as the production of HO, the possible mechanism for ozonation of BA at acidic pH was proposed.
2.
Materials and methods
2.1.
Reagents
Benzoic acid (BA, 99.5%), 2-hydroxybenzoic acid (2-HBA, 99.5%), 3-hydroxybenzoic acid (3-HBA, 99%), 4-hydroxybenzoic acid (4-HBA, 99%) were supplied by Sigma-Aldrich. Other reagents were at least of analytical grade. All the solutions in this study were prepared with milli-Q water (18.2 MUcm).
2.2.
Experimental procedure
Ozone gas was generated by an oxygen-fed ozonator (CF-G-250 g, QingDao Guolin, China), and its stock solutions were produced and thermostated at 4 C. Batch ozonation experiments were performed in the borosilicate glass apparatus at 25 C (thermostated with water bath) to exclude UV contribution to HO formation. The reactions started by spiking ozone stock solution into solution buffered by 10 mM phosphate. Sulfite was used to quench the residual ozone in the samples. All the experiments were repeated at least three times.
2.3.
Methods
The pH was measured by a pH meter (PB-10, Sartorious). Ozone concentrations were determined spectrophotometrically on UVevis spectrophotometer directly by using adsorption coefficient (ε) of 3000 M1 cm1 of ozone at 258 nm or by , 1981). H2O2 conusing the indigo method (Bader and Hoigne centration was measured by DPD/POD method (Bader et al., 1988). BA and HBAs were analyzed on HPLC (Ultimate 3000, Dionex) equipped with a UV detector, using a C18 column (4.6 mm 100 mm, 3.5 mm particle size, Agilent Eclipse Plus) and an eluent consisting of phosphoric acid solution (10 mM) and methanol (55:45, v/v%) at a flow rate of 1.0 mL/min. The concentrations of BA and HBA were quantified at the wavelength of 227 nm and 250 nm, respectively.
w a t e r r e s e a r c h 7 3 ( 2 0 1 5 ) 9 e1 6
2.4.
Electron paramagnetic resonance (EPR)
EPR experiments were conducted on a Bruker EMX 10/12 spectrometer (Germany) at room temperature. Its basic parameters include a resonance frequency of 9.77 GHz, microwave power of 20.02 mW, modulation frequency of 100 kHz, modulation amplitude of 1.0 G, sweep width of 100 G, time constant of 40.96 ms, sweep time of 83.89 s, and receiver gain 1.0 103. The solution contained 10 mM HBA, 100 mM DMPO (pH 7.4, 10 mM phosphate buffer) and 100 mM ozone at pH 2.30.
3.
Results and discussions
3.1.
BA degradation by ozone at pH 2.3
3.1.1.
BA degradation
The degradation efficiency of BA and ozone decomposition during ozonation at pH 2.3 was shown in Fig. 1. Efficient degradation of BA was observed. About 85% of BA was degraded within 10 min, and the degradation kinetics of BA followed pseudo first order. It is known that two oxidizing species, i.e. molecular ozone and hydroxyl radicals (HO), are generally involved in ozonation. However, ozone reacts very and Bader, 1983), and slowly with BA (k < 1.2 M1s1) (Hoigne the carboxylic group of BA is expected to deactivate the electrophilic aromatic substitution reaction with ozone. Actually, BA is usually used as the HO probe in many AOPs (Chen et al., 2011). Thus, one can assume that BA degradation during ozonation at pH 2.3 is a HO involved reaction, as further discussed later.
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a well-known HO scavenger, on ozone decomposition and BA degradation. As shown in Fig. 2, TBA greatly or even completely inhibited the ozone decomposition and BA degradation, suggesting that ozone is decomposed through a HO radical reaction, and BA is degraded through reaction with HO. Thus, one can speculate the intermediates formed during BA degradation were capable of activating ozone to decompose with production of HO.
3.2.
Effect of BA and ozone concentrations
3.2.1.
Effect of BA concentration on the degradation
Fig. 3a shows the degradation efficiency of BA at initial concentrations from 0.2 to 10 mM and pH 2.3. It was considerably enhanced by increasing the initial BA concentration from 0.2 to 5 mM, and then dropped to 10 mM. At the time of 10 min, the degradation efficiency of BA was about just 20% for 0.2 mM BA. It was increased to about 85% for 5.0 mM BA, followed by a gradual decrease for higher concentration of BA. It is interesting that the degradation of BA showed significant and inverted parabolic dependency on the concentration of target compounds. It is quite different from most degradation reactions, with their efficiency increasing with the decreasing concentration (Hu et al., 2008). Besides, the initial degradation rate of BA was also promoted with the increase of BA concentration (Fig. S1). Thus, it could be assumed that BA degradation by ozone at pH ¼ 2.3 is a self-enhanced reaction.
3.2.2.
Effect of BA concentration on ozone decomposition
Concurrently, ozone decomposition was also enhanced significantly in the presence of 3 mM BA (Fig. 1). More than 60% of ozone was decomposed within 10 min, whereas less than 15% in the absence of BA. The simultaneous rapid degradation of BA and decomposition of ozone suggested that BA induced ozone decomposition resulted in the production of HO. To verify this point, we examined the effect of tert-butanol (TBA),
As shown in Fig. 3b, enhanced ozone decomposition was observed by increasing the BA concentration from 0 to 10 mM. At the time of 10 min, less than 20% of ozone was decomposed for 0.2 mM BA, close to self-decomposition of ozone, whereas 30e60% of ozone was decomposed by increasing the initial BA concentration from 0.5 to 2 mM. Further increase in BA concentration would result in the complete decay of ozone. The kinetic of ozone decomposition at the different BA concentrations followed the first order apparently. In addition, increasing BA concentration significantly would accelerate the initial decomposition rate of ozone (Fig. S1). All these
Fig. 1 e BA degradation and ozone decomposition under acidic conditions. Condition: pH 2.3 ± 0.1, 25 C, [O3]0 ¼ 1.0 ± 0.05 mg/L, and [BA]0 ¼ 3 mM.
Fig. 2 e Influence of TBA on BA degradation and ozone decomposition. Condition: pH 2.3 ± 0.1, 25 C, [O3]0 ¼ 1.0 ± 0.05 mg/L, [BA]0 ¼ 10 mM, and [TBA]0 ¼ 10 mM.
3.1.2.
Ozone decomposition
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results further indicated that the intermediates formed during BA degradation was capable of promoting ozone decomposition to produce HO.
3.2.3.
Effect of ozone dosage on BA degradation
Generally, ozone dose is a key parameter in ozonation process. Fig. 4 shows the degradation of BA with various doses of ozone (0.5, 1.0, 2.0, and 4.0 mg/L). The increase in ozone concentration from 0.5 mg/L to 4.0 mg/L increased the BA degradation efficiency from 60% to 95%. It is generally recognized that the ozone exposure mainly influences the degradation of the ozone-reactive pollutants (kO3 > 103 M1s1) and has little impact on that of the ozone-refractory pollutants (kO3 < 10 M1s1) (Buffle et al., 2006). However, in this study the ozone dose plays an important role in the degradation efficiency of the ozone-refractory HO probe, BA. The possible reason is that the increasing ozone dose accelerated the formation of intermediate products of BA for activating ozone to yield more HO, as elucidated below. Generally, aromatic organic compounds substituted with electron-withdrawing groups (such as eCOOH, eNO2, and eCl) are recalcitrant to direct attack of ozone. Also, we generally believe that those aromatics are refractory by ozonation at acidic pHs, where little HO would be generated due to the lack of OH. Unexpectedly, the results reported herein indicated that that ozonation may be used in effective removal of recalcitrant aromatic compounds from acidic wastewater.
3.3.
Mechanism
3.3.1.
Effect of impurities
It has been reported that OH, H2O2, electron-rich compounds (e.g, phenols, amines, alkoxylated aromatics), transition metal ions, and some metal oxides can initiate ozone to produce HO (Staehelin et al., 1982; Buffle et al., 2006; Mvula et al., 2009; Kasprzyk-Horden et al., 2003). It is noted that all solutions in this study were prepared with milli-Q water (18 MUcm) and that all the reagents are at least of ACS reagent grade. The concentration of metal ions was below the detection limit in solution by using AAS equipped with graphite furnace. Hence, the effect of such possible impurities in the reagents on ozone decomposition and HO formation should be excluded. The pre-ozonation of the milli-Q solution showed its negligible effect on the degradation efficiency of BA, also excluding the background contribution to HO formation (Fig. S2).
3.3.2.
Role of H2O2
The solution was buffered at pH 2.3 using 10 mM H3PO4, where [OH] is about 2.0 1012 M. The apparent rate constant of ozone decomposition initiated by OH is calculated to be , 1982), which is so low 3.5 1010 s1 (Staehelin and Hoigne that the role of OH in initiation of ozone could be neglected. It has been reported that H2O2 and its derivatives (e.g., HO2 ) intermediately formed during ozonation of organics influence the ozone stability in aqueous solutions (Staehelin et al., 1985; Pi et al., 2005) at neutral pH. At acidic pHs like 2.3, less than 0.3 mM H2O2 is formed (Fig. S3a), which is insufficient to induce ozone decay to produce HO, considering the low level of HO2 (pKa of ~4.8) and the slow rate constant of reaction ozone with , 1982). It H2O2 (kO3 ;H2 O2 < 0.01 M1s1) (Staehelin and Hoigne was further confirmed by the slight ozone decomposition with the addition of 10 or 20 mM H2O2 (Fig. S3b). Thus, we believe that BA degradation does not rely on the production of H2O2 under acidic conditions.
3.3.3.
Role of hydroxylated intermediates
One typical process of ozone reaction with aromatic compounds is hydroxylation. It has been proved that phenolic intermediates are responsible for promoting HO production € the et al., 2009). However, it is difficult to directly detect the (No
Fig. 3 e (a) BA degradation and (b) ozone decomposition at different initial BA concentrations. Condition: pH 2.3 ± 0.1, 25 C, [O3]0 ¼ 1.0 ± 0.05 mg/L ※ blank - 0.2 mM B 0.5 mM : 1.0 mM 7 2.0 mM = 5.0 mM 10.0 mM.
Fig. 4 e BA degradation at different ozone dosage. Condition: pH 2.3 ± 0.1, 25 C, and [BA]0 ¼ 2.0 mM.
w a t e r r e s e a r c h 7 3 ( 2 0 1 5 ) 9 e1 6
hydroxylated intermediates because they react with ozone much faster than the parent aromatics, causing their rapid consumption by ozone. For instance, the second-order rate constant for ozone reaction with 4-hydroxylbenzoic acid (4HBA) at pH 2 was 1.8 104 M1 s1 (Benitez et al., 2000). Unluckily, we cannot obtain the rate constant values for 2-HBA and 3-HBA reacting with ozone. In fact, no 2- and 3-HBAs but trace 4-HBA was detected during ozonation of 100 mM BA, as shown in Fig. S4. We assume that, if all the three HBAs are intermediately formed, the reaction of ozone with 2- and 3-HBAs should be more rapid than 4-HBA. In fact, as shown in Fig. S5, the sequence of ozone decomposition initiated by three HBAs followed 2-HBA > 3- HBA > 4-HBA. Fig. 5 shows the effect of 10 mM 2-, 3-, and 4-HBA on the degradation of BA. It was substantially accelerated within 1 min and then stopped because ozone was entirely exhausted. Apparently, all the three HBAs could increase HO exposure for BA degradation. The significant increase in HO production by the presence of HBAs was further verified by EPR study (Fig. 6), suggesting that hyroxybenzoic acids could induce ozone decomposition to produce HO at acidic pHs. Fig. S5 shows different inhibitory effect of TBA on ozone decomposition in the presence of various HBAs and phenol. TBA exhibited strong inhibition on the ozone decomposition involved by 3- and 4-HBA, moderate for 2-HBA and only mild for phenol, suggesting that a substantial fraction of ozone participated in the radical reactions, especially for 3-HBA and 4-HBA. This means that HO generation induced by hyroxybenzoic acid is similar to that by phenol (Mvula et al., 2003;
Fig. 5 e Influence of HBA on BA degradation and change of HBA during ozonation. Condition: pH 2.3 ± 0.1, 25 C, [O3]0 ¼ 1.0 ± 0.05 mg/L, [BA]0 ¼ 2 mM, [HBA]0 ¼ 10 mM, and [TBA]0 ¼ 2.5 mM.
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Buffle et al., 2006). Upon the ozonation of phenols, HO is produced through an electron transfer from phenolate/phenol to ozone, and the HO yield of phenol was about 22% in a wide pH range from 3 to 10 (Mvula and von Sonntag, 2003). The fact that TBA decreases ozone decomposition in the presence of HBAs and phenol (Fig.S5) suggests a similar mechanism for HBAs reaction with ozone as phenol, i.e., via an electron transfer mechanism. In the hydroxylation of BA, the molar ratio for formation of 2-, 3- and 4-HBA is about 1.4:2:1 (Klein et al., 1975). Their corresponding stoichiometric ratio of reacting with ozone was about 1:1.77, 1:2.57, 1:4.16, respectively (Fig. S6). Thus, we can calculated that 0.29 mol HBA could be destroyed by per mol of ozone consumed, which is close to the stoichiometry of ozone reaction with phenol (0.33) (Mvula and von Sonntag, 2003). From the above analysis, we believe that the secondary reactions of three HBAs with ozone should be involved in the HO formation and ozone decomposition.
3.3.4.
Possible reaction pathways
Based on the above analysis and the available references on ozonation, we suggested BA ozonation at acidic pHs possibly undergoes the following reactions:
BA reacts rapidly with HO (up to ~109 M1s1) to give four isomeric HO-adduct radicals (reaction 1). Similar results were reported on hydroxylation of phenol (Klein et al., 1975; Mvula et al., 2001). The first production of HO in reaction 1 is likely to derive from instantaneous decomposition of ozone in the initial phase observed in Fig. 3b, as supported by a weak DMPO-OH signal during the decomposition of ozone alone at pH 2.3 (Fig. 6). Till now several factors have been propose for the initial production of HO, such as thermal decomposition of ozone, direct ozone-water reaction, bubble breakage in agitation process and glass container (Sotelo et al., 1987; Sehested et al., 1991; 1998; Jin et al., 2012). The transient
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Fig. 6 e EPR spectra of reaction between ozone with HBA. Condition: pH 2.3 ± 0.1, 25 C, [O3]0 ¼ 5.0 mg/L, [HBA]0 ¼ 10 mM, [DMPO]0 ¼ 100 mM a: DMPO; b: DMPO þ 2HBA; c: DMPO þ O3; d: DMPO þ O3 þ 2HBA; e: DMPO þ O3 þ 3HBA; f: DMPO þ O3 þ 4HBA.
HO-adduct radicals produced during hydroxylation are highly reactive with oxidants like O2, and O3 (von Sonntag, 1991, Von Sonntag, 2006). Thus, in O3/O2 aqueous solution, HO-adduct radicals would involve in the competition of O3 and O2 (reaction 2 and 4). To examine the role of O2 in BA degradation, we performed BA degradation and ozone decomposition experiments under N2-sparged (2.0 105 M O2), O2-sparged (1.0 103 M O2) and raw (2.5 104 M O2) conditions, and the results are shown in Fig. 7. At the time of 10 min, even the increase of 50-time oxygen concentration from N2-sparged to O2-sparged conditions just had mild effect on BA degradation and ozone decomposition. It suggested that HO-adduct radical reaction with ozone is dominant, even though O2 is at least in a 10-fold excess over O3 in the experimental solutions. It is reasonable because of the very low rate of HO-adduct radicals substituted with eCOOH reacting with O2 (the second-order rate constant ~106 M1s1, Fang et al., 1995). Naumov and von Sonntag (2011) suggested that ozone reacted with HO-adduct radicals by oxygen transfer based on
quantum-chemical calculations, and no HO, O3 $ or O2 $ /HO2 is generated like as during ozonation of sulfurcontaining compounds. It is not the case for ozone decomposition through a HO chain reaction during ozonation of BA. We think that ozone is unlikely to react with HO-adducts by oxygen transfer, and electron transfer seems to make sense. The HO-adduct radicals substituted by eNO2, eCN, eOCH3 and eF are oxidized to phenolic compounds by Fe3þ, Cu2þ, Cr6þ through electron-transfer (Eberhardt, 1977). Ozone has stronger oxidative ability and would be likely to react with the isomeric HO-adduct radicals by electron transfer to yield O3 $ , which is a necessary and effective precursor of HO via reaction 6 and 7. The hydroxylation species are further oxidized by O3 to produce O3 $ via reaction 3. With regard to the reactivity with oxygen, each of HOadduct radicals may in turn yield two different peroxyl radicals, namely, the 1,3- and 1,4-cyclohexadienyl-type species (Fang et al., 1995), whereby 1,3-cyclohexadienyl-type specie is only the minor component. It has been observed that rapid HO2 formation can only result from the peroxyl radicals of the cyclohexa-1,3-diene structure but not that for the cyclohexa1,4-diene structure (Pan et al., 1993). Hence, only a minor part of peroxyl radicals can eliminate HO2∙ and thereafter makes a slight contribution to HO formation via reaction 5 and 8. Thus, the pathway of electron transfer between O3 and HO-adduct radicals contributed the main source of HO, while HO2 accounts for only minor source of HO. Based on eq. (2e3), 1 mol BA would consume 4 mol of ozone in theory, i.e., 1 mol by HO-adduct radicals and 3 mol of by formed hydroxybenzoic acid via electron transfer. Fig. 2 shows that 20 mM ozone was completely decayed, whereas about 5 mM BA was degraded within 10 min. The observed stoichiometric ratio of about 1:4 for BA degradation and ozone decomposition is consistent with both equations. The intermediates produced during ozonation of BA have been shown to be capable of activating ozone to generate HO. Actually, the intermediates formed during ozonation of aromatics substituted with other EWGs such as eNO2 and eCl are also able to initiate ozone to produce HO in our study (seen in Supporting Information for review only), and those findings will be reported later. In summary, it is a common property for the aromatic compounds substituted by electronwithdrawing group such as eCN, eF, eBr, eI, eCHO, eCOOR, eNH3 þ and eSO3 . It should be noted that the proposed mechanism is not fully justified. To better understand the mechanism, further studies are still required to identify the reaction intermediate products and pathways of HO production. In addition, the roles of different electron-withdrawing groups in HO yield should be further examined.
4.
Conclusions
Based on this study, we may make the following conclusions:
Fig. 7 e Effect of oxygen on BA degradation and ozone decomposition. Condition: pH 2.3 ± 0.1, 25 C, [O3]0 ¼ 1.0 ± 0.05 mg/L, [BA]0 ¼ 10 mM.
(1) Benzoic acid was degraded efficiently by ozone via HOinduced mechanism under acid conditions, and the efficiency was significantly enhanced by the increased concentration of BA and ozone dosage.
w a t e r r e s e a r c h 7 3 ( 2 0 1 5 ) 9 e1 6
(2) The intermediates produced during BA degradation are involved in ozone activation to generate HO as well as ozone decomposition. (3) Under acidic conditions, the precursor of HO, i.e. O3 $ was formed mainly through the electron transfer reaction of ozone with HO-adduct radicals accounted for the major contributions, while HO2 elimination of peroxyl radicals make minor contributions (4) Hydroxylated intermediates played a positive role in the HO production, as evidenced by EPR study and products analysis.
Acknowledgment This research was supported by the Program of National Natural Science Foundation of China (21307057), the Program of Natural Science Foundation of Jiangsu Province (BK20130577), the Specialized Research Fund for the Doctoral Program of Higher Education of China (SRFDP, 20130091120014), and the Fundamental Research Funds for the Central Universities (20620140128).
Appendix A. Supplementary data Supplementary data related to this article can be found at http://dx.doi.org/10.1016/j.watres.2015.01.010.
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Fang, X.M., Pan, X.M., Rahmann, A., Schumann, H.P., von Sonntag, C., 1995. Reversibility in the reaction of cyclohexadienyl radicals with oxygen in aqueous solution. Chem. Eur. J. 1 (7), 423e429. Gogate, P.R., Pandit, A.B., 2004. A review of imperative technologies for wastewater treatment I: oxidation technologies at ambient conditions. Adv. Environ. Res. 8 (3e4), 501e551. , J., Bader, H., 1983. Rate constants of reactions of Hoigne ozone with organic and inorganic compounds in water. I. Non-dissociating organic compounds. Water Res. 17 (2), 173e183. Hu, C., Xing, S.T., Qu, J.H., He, H., 2008. Catalytic ozonation of herbicide 2,4-D over cobalt oxide supported on mesoporous zirconia. J. Phys. Chem. C 112 (15), 5978e5983. Ikehata, K., El-Din, M.G., Snyder, S.A., 2008. Ozonation and advanced oxidation treatment of emerging organic pollutants in water and wastewater. Ozone Sci. Eng. 30 (1), 21e26. Jin, B., Su, M.N., Lin, J.J.M., 2012. Does ozone-water complex produce additional OH radicals in the atomosphere? J. Phys. Chem. A 116 (49), 12082e12088. łek, M., Nawocki, J., 2003. Catalytic Kasprzyk-Horden, B., Zio ozonation and methods of enhancing molecular ozone reactions in water treatment. Appl. Catal. B: Environ. 46 (4), 639e669. Kim, K.H., Ihm, S.K., 2011. Heterogeneous catalytic wet air oxidation of refractory organic pollutants in industrial wastewaters: a review. J. Hazard. Mater. 186 (1), 16e34. Klein, G.W., Bhatla, K., Madhavan, V., Schuler, R.H., 1975. Reaction of OH with benzoic acid. Isomer distribution in the radical intermediates. J. Phys. Chem. 79 (17), 1767e1774. Marcucci, M., Nosenzo, G., Capannelli, G., Ciabatti, I., Corrieri, D., Ciardelli, G., 2001. Treatment and reuse of textile effluents based on new ultrafiltration and other membrane technologies. Desalination 138 (1e3), 75e82. Merga, G., Schuchmann, H.P., Madhava Rao, B.S., von Sonntag, C., 1996. OH radical-induced oxidation of chlorobenzene in aqueous solution in the absence and presence of oxygen. J. Chem. Soc. Perkin Trans. 2, 1097e1103. Mvula, E., Naumov, S., von Sonntag, C., 2009. Ozonolysis of lignin models in aqueous solution: anisole, 1,2-dimethoxybenzene, 1,4-dimethoxybenzene, and 1,3,5-trimethoxybenzene. Environ. Sci. Technol. 43 (16), 6275e6282. Mvula, E., Schuchmann, M.N., von Sonntag, C., 2001. Reactions of phenol-OH-adduct radicals. Phenoxyl radical formation by water elimination vs. oxidation by dioxygen. J. Chem. Soc. Perkin Trans. 2, 264e268. Mvula, E., von Sonntag, C., 2003. Ozonolysis of phenols in aqueous solution. Org. Biomol. Chem. 1, 1749e1756. Naumov, S., von Sonntag, C., 2011. Standard gibbs free energies of reactions of ozone with free radicals in aqueous solution: quantum-chemical calculations. Environ. Sci. Technol. 45 (21), 9195e9204. € the, T., Fahlenkamp, H., von Sonntag, C., 2009. Ozonation of No wastewater: rate of ozone consumption and hydroxyl radical yield. Environ. Sci. Technol. 43 (15), 5990e5995. nchez-Pe rez, J.A., 2011. Combination of Oller, I., Malato, S., Sa advanced oxidation processes and biological treatments for wastewater decontaminationdA review. Sci. Total Environ. 409 (20), 4141e4166. Pan, X.M., Schuchmann, M.N., von Sonntag, C., 1993. Oxidation of benzene by the OH radical. a product and pulse radiolysis study in oxygenated aqueous solution. J. Chem. Soc. Perkin Trans. 2, 289e297. Panizza, M., Bocca, C., Cerisola, G., 2000. Electrochemical treatment of wastewater containing polyaromatic organic pollutants. Water Res. 34 (9), 2601e2605.
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Pham, A.L., Lee, C., Doyle, F.M., Sedlak, D.L., 2009. A silicasupported iron oxide catalyst capable of activating hydrogen peroxide at neutral pH values. Environ. Sci. Technol. 43 (23), 8930e8935. Pi, Y.Z., Schumacher, J., Jekel, M., 2005. Decomposition of aqueous ozone in the presence of aromatic organic solutes. Water Res. 39 (1), 83e88. Raghavan, N.V., Steenken, S., 1980. Electrophilic reaction of the hydroxyl radical with phenol. Determination of the distribution of isomeric dihydroxycyclohexadienyl radicals of the distribution of isomeric dihydroxycyclohexadienyl radicals. J. Am. Chem. Soc. 102 (10), 3495e3499. Sehested, K., Corfitzen, H., Holcman, J., Fischer, C.H., Hart, E.J., 1991. The primary reaction in the decomposition of ozone in acidic aqueous solution. Environ. Sci. Technol. 25 (9), 3057e3063, 1589-1596. Sehested, K., Corfitzen, H., Holcman, J., Fischer, C.H., Hart, E.J., 1998. On the mechanism of the decomposition of acidic O3 solutions, thermally or H2O2-initiated. J. Phys. Chem. A 102 (16), 2267e2672. Sirtori, C., Zapata, A., Oller, I., Gernjak, W., Aguera, A., Malato, S., 2009. Decontamination industrial pharmaceutical wastewater by combining solar photo-Fenton and biological treatment. Wat. Res. 43 (3), 661e668. n, F.J., Benı´tez, F.J., Beltra n-Heredia, J., 1987. Sotelo, J., Beltra Ozone decomposition in water: kinetic study. Ind. Eng. Chem. Res. 26 (1), 39e43.
, J., 1982. Decomposition of ozone in water: Staehelin, J., Hoigne rate of initiation by hydroxide ions and hydrogen peroxide. Environ. Sci. Technol. 16 (10), 676e681. , J., 1985. Decomposition of ozone in water in Staehelin, J., Hoigne the presence of organic solutes acting as promoters and inhibitors of radical chain reactions. Environ. Sci. Technol. 19 (12), 1206e1213. Steenken, S., Raghavan, N.V., 1979. Oxidation of methoxyhydroxycyclohexadienyl radicals by quinones. The influence of the isomeric structure of the radical on the rate constant for electron transfer. J. Phys. Chem. 83 (24), 3101e3107. Von Gunten, U., 2003. Ozonation of drinking water: Part I. Oxidation kinetics and product formation. Water Res. 37 (5), 1443e1467. Von Sonntag, C., 2006. Free Radical Induced DNA Damage and its Repair a Chemical Perspective. Springer, Berlin. Von Sonntag, C., Schuchmann, H.P., 1991. The elucidation of peroxyl radical reactions in aqueous solution with the help of radiation-chemical methods. Angew. Chem. Int. Ed. Engl. 30 (10), 1229e1253. Yang, W.B., Zhou, H.D., Cicek, N., 2014. Treatment of organic micropollutants in water and wastewater by UV-based processes: a literature review. Crit. Rev. Env. Sci. Technol. 44 (13), 1443e1476.
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Self-enhanced ozonation of benzoic acid at acidic pHs Xianfeng Huang a, Xuchun Li a,b, Bingcai Pan a,*, Hongchao Li a, Yanyang Zhang a, Bihuang Xie a a
State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing 210023, China b School of Environmental Science and Engineering, Zhejiang Gongshang University, Hangzhou 310012, China
article info
abstract
Article history:
Ozonation of recalcitrant contaminants under acidic conditions is inefficient due to the
Received 2 September 2014
lack of initiator (e.g., OH) for ozone to produce hydroxyl radicals (HO). In this study, we
Received in revised form
reported that benzoic acid (BA), which is inert to ozone attack, underwent efficient
2 January 2015
degradation by ozone at acidic pH (2.3). The kinetics of BA degradation and ozone
Accepted 7 January 2015
decomposition were both enhanced by increasing BA concentrations. Essentially, it is a
Available online 16 January 2015
HO-mediated reaction. Based on the exclusion of possible contributions of H2O2 and
Keywords:
mation of ozone ion (O3 $ ), which is an effective precursor of HO, was thus proposed. The
Ozone
hydroxycyclohexadienyl-type radicals generated during the attack of BA by HO may lead
Acidic conditions
to the formation of O3 $ . Meanwhile, O3 $ could also be possibly formed from the reaction
Benzoic acid
between ozone and organic (e.g., ROO∙) or inorganic peroxyl radicals (e.g., HO2). In addi-
phenol-like intermediates for HO production, the reaction mechanism involved the for-
tion, the hydroxylated products like phenol-like intermediates also played a positive role in
HO
Advanced oxidation process
HO production. Consequently, HO was produced efficiently under acidic conditions, resulting in rapid degradation of BA. This study provides a new approach for ozone activation even at acidic pHs, and broadens the knowledge of ozonation in removal of micropollutants from water. © 2015 Elsevier Ltd. All rights reserved.
1.
Introduction
Nowadays, an increasing discharge of acidic wastewater from various industries, e.g., pesticide, printing and dyeing, pharmacy, chemical industry and electroplating, shows potential risks to ecosystem and human health (Gogate et al., 2004; Oller et al., 2011). A variety of organic compounds therein, including pesticides, pharmaceuticals, PCBs, PAHs, nitroaromatics, and dyes, need advanced treatment due to their carcinogenic, * Corresponding author. Tel.: þ86 25 8968 0390. E-mail address: [email protected] (B. Pan). http://dx.doi.org/10.1016/j.watres.2015.01.010 0043-1354/© 2015 Elsevier Ltd. All rights reserved.
teratogenic and mutagenic effects (Ikehata et al., 2008; Sirtori et al., 2009; Kim et al., 2011). However, efficient and safe removal of these contaminants from acidic wastewater is still challenging now. Currently, several methods have been developed and widely used to treat acidic wastewater, such as biological process, membrane separation, and advanced oxidation (AOPs) (Oller et al., 2011; Marcucci et al., 2001; Gogate et al., 2004). The hydroxyl radical (HO)-involved AOPs, including
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Fenton oxidation, wet oxidation, photochemical oxidation and electrochemical oxidation, are very promising and have been attracting great interests. Despite their different advantages in water treatment, some inherent disadvantages have to face now. In brief, Fenton oxidation has to face the high reagent dosage, iron sludge yield and hydrogen peroxide residue (Pham et al., 2009). Wet oxidation is usually performed under high temperature and pressure. Thus, the reaction equipment should be of resistance to high temperature/pressure and corrosion, resulting in high investment and operation cost (Kim and Ihm, 2011). In addition, most of the photochemical oxidation is still far from real application because it is inevitably restricted to short light penetration distance in wastewater and high energy cost (Yang et al., 2014). Also, electrochemical oxidation still encounters some problems such as the electrode corrosion and passivation and low current efficiency (Panizza et al., 2000). In general, further improvement is still required to forward the above AOPs to wider application in wastewater treatment. Comparatively, ozonation is widely used for taste and odor control, decoloration, elimination of organic and inorganic contaminants, by direct oxidation using molecular ozone and/or through radical reactions involving HO (Von Gunten, 2003). The HO mediated ozonation is attractive and promising in removal of recalcitrant contaminants, such as nitrobenzene, pesticides, pharmaceuticals, and even persistent organic pollutants (Ikehata et al., 2008; Buffle et al., 2006). However, the production of HO from ozone needs initiator like hydroxide ions (OH). Thus, its efficiency is dramatically decreased under acidic conditions. To improve the efficiency of HO generation during ozonation at acidic pHs, some effective strategies have been proposed, including combination of H2O2, energy input (e.g., UV and ultrasound), and addition of transition metal ions or transition metal oxides (Kasprzyk-Horden et al., 2003). Recently, some electron-rich moieties (ERMs) such as phenols, amines, alkoxybenzenes, and hydroxylation product formed intermediately during ozonation have been identified to be capable of activating O3 to produce HO radicals (Buffle et al., 2006b; Mvula et al., 2009). On the contrary, little attention was paid to O3 activation by electron-deficient moieties (EDMs) such as aromatics compounds substituted by electron-withdrawing groups (EWG) or their intermediate species. This is possibly because they are usually ozonerefractory. On the other hand, compounds substituted with EWG such as eCOOH, eNO2, eCl, and eCN could be efficiently oxidized by HO radicals at second-order rate constants from ~108 to ~1010 M1 s1, accompanying the formation of hydroxycyclohexadienyl-type radicals (HO-adduct radicals) (Merga et al., 1996; von Sonntag, 2006). Those HO-adduct radicals can serve as oxygen-transfer media or electron shuttle (Von Sonntag, 2006; Pan et al., 1993; Duesterberg et al., 2007), and thus react with a variety of oxidants such as O2, quinone, and metal ions (e.g., Fe3þ, Cu2þ, Cr6þ) (Steenken et al., 1979; Raghavan et al., 1980). The HO-adduct radicals also react irreversibly with O3 via oxygen transfer based on quantum-chemical calculation (Naumov and von Sonntag, 2011). Considering much stronger oxidizing ability of O3
than the above oxidants, the HO-adduct radicals may participate in the formation of ozone ion (O3 $ ), a precursor of HO, and the enhanced production of HO by those radicals in ozonation under acidic conditions could be anticipated. However, the efficiency of ozone activation by aromatics substituted with EWG or their intermediates at acidic pHs is still unclear. The primary objective of this study was to investigate the efficiency of ozone activation by aromatics substituted with EWG or their intermediates at acidic pHs, benzoic acid being employed as a representative compound. Kinetics of BA degradation and ozone decomposition at acidic pH (2.3) were investigated. Effects of the concentration of BA and ozone dosage on the HO production efficiency were evaluated. Electron paramagnetic resonance studies (EPR) was used to identify the production of HO. By analyzing the roles of reactive species intermediately generated in the activation of ozone as well as the production of HO, the possible mechanism for ozonation of BA at acidic pH was proposed.
2.
Materials and methods
2.1.
Reagents
Benzoic acid (BA, 99.5%), 2-hydroxybenzoic acid (2-HBA, 99.5%), 3-hydroxybenzoic acid (3-HBA, 99%), 4-hydroxybenzoic acid (4-HBA, 99%) were supplied by Sigma-Aldrich. Other reagents were at least of analytical grade. All the solutions in this study were prepared with milli-Q water (18.2 MUcm).
2.2.
Experimental procedure
Ozone gas was generated by an oxygen-fed ozonator (CF-G-250 g, QingDao Guolin, China), and its stock solutions were produced and thermostated at 4 C. Batch ozonation experiments were performed in the borosilicate glass apparatus at 25 C (thermostated with water bath) to exclude UV contribution to HO formation. The reactions started by spiking ozone stock solution into solution buffered by 10 mM phosphate. Sulfite was used to quench the residual ozone in the samples. All the experiments were repeated at least three times.
2.3.
Methods
The pH was measured by a pH meter (PB-10, Sartorious). Ozone concentrations were determined spectrophotometrically on UVevis spectrophotometer directly by using adsorption coefficient (ε) of 3000 M1 cm1 of ozone at 258 nm or by , 1981). H2O2 conusing the indigo method (Bader and Hoigne centration was measured by DPD/POD method (Bader et al., 1988). BA and HBAs were analyzed on HPLC (Ultimate 3000, Dionex) equipped with a UV detector, using a C18 column (4.6 mm 100 mm, 3.5 mm particle size, Agilent Eclipse Plus) and an eluent consisting of phosphoric acid solution (10 mM) and methanol (55:45, v/v%) at a flow rate of 1.0 mL/min. The concentrations of BA and HBA were quantified at the wavelength of 227 nm and 250 nm, respectively.
w a t e r r e s e a r c h 7 3 ( 2 0 1 5 ) 9 e1 6
2.4.
Electron paramagnetic resonance (EPR)
EPR experiments were conducted on a Bruker EMX 10/12 spectrometer (Germany) at room temperature. Its basic parameters include a resonance frequency of 9.77 GHz, microwave power of 20.02 mW, modulation frequency of 100 kHz, modulation amplitude of 1.0 G, sweep width of 100 G, time constant of 40.96 ms, sweep time of 83.89 s, and receiver gain 1.0 103. The solution contained 10 mM HBA, 100 mM DMPO (pH 7.4, 10 mM phosphate buffer) and 100 mM ozone at pH 2.30.
3.
Results and discussions
3.1.
BA degradation by ozone at pH 2.3
3.1.1.
BA degradation
The degradation efficiency of BA and ozone decomposition during ozonation at pH 2.3 was shown in Fig. 1. Efficient degradation of BA was observed. About 85% of BA was degraded within 10 min, and the degradation kinetics of BA followed pseudo first order. It is known that two oxidizing species, i.e. molecular ozone and hydroxyl radicals (HO), are generally involved in ozonation. However, ozone reacts very and Bader, 1983), and slowly with BA (k < 1.2 M1s1) (Hoigne the carboxylic group of BA is expected to deactivate the electrophilic aromatic substitution reaction with ozone. Actually, BA is usually used as the HO probe in many AOPs (Chen et al., 2011). Thus, one can assume that BA degradation during ozonation at pH 2.3 is a HO involved reaction, as further discussed later.
11
a well-known HO scavenger, on ozone decomposition and BA degradation. As shown in Fig. 2, TBA greatly or even completely inhibited the ozone decomposition and BA degradation, suggesting that ozone is decomposed through a HO radical reaction, and BA is degraded through reaction with HO. Thus, one can speculate the intermediates formed during BA degradation were capable of activating ozone to decompose with production of HO.
3.2.
Effect of BA and ozone concentrations
3.2.1.
Effect of BA concentration on the degradation
Fig. 3a shows the degradation efficiency of BA at initial concentrations from 0.2 to 10 mM and pH 2.3. It was considerably enhanced by increasing the initial BA concentration from 0.2 to 5 mM, and then dropped to 10 mM. At the time of 10 min, the degradation efficiency of BA was about just 20% for 0.2 mM BA. It was increased to about 85% for 5.0 mM BA, followed by a gradual decrease for higher concentration of BA. It is interesting that the degradation of BA showed significant and inverted parabolic dependency on the concentration of target compounds. It is quite different from most degradation reactions, with their efficiency increasing with the decreasing concentration (Hu et al., 2008). Besides, the initial degradation rate of BA was also promoted with the increase of BA concentration (Fig. S1). Thus, it could be assumed that BA degradation by ozone at pH ¼ 2.3 is a self-enhanced reaction.
3.2.2.
Effect of BA concentration on ozone decomposition
Concurrently, ozone decomposition was also enhanced significantly in the presence of 3 mM BA (Fig. 1). More than 60% of ozone was decomposed within 10 min, whereas less than 15% in the absence of BA. The simultaneous rapid degradation of BA and decomposition of ozone suggested that BA induced ozone decomposition resulted in the production of HO. To verify this point, we examined the effect of tert-butanol (TBA),
As shown in Fig. 3b, enhanced ozone decomposition was observed by increasing the BA concentration from 0 to 10 mM. At the time of 10 min, less than 20% of ozone was decomposed for 0.2 mM BA, close to self-decomposition of ozone, whereas 30e60% of ozone was decomposed by increasing the initial BA concentration from 0.5 to 2 mM. Further increase in BA concentration would result in the complete decay of ozone. The kinetic of ozone decomposition at the different BA concentrations followed the first order apparently. In addition, increasing BA concentration significantly would accelerate the initial decomposition rate of ozone (Fig. S1). All these
Fig. 1 e BA degradation and ozone decomposition under acidic conditions. Condition: pH 2.3 ± 0.1, 25 C, [O3]0 ¼ 1.0 ± 0.05 mg/L, and [BA]0 ¼ 3 mM.
Fig. 2 e Influence of TBA on BA degradation and ozone decomposition. Condition: pH 2.3 ± 0.1, 25 C, [O3]0 ¼ 1.0 ± 0.05 mg/L, [BA]0 ¼ 10 mM, and [TBA]0 ¼ 10 mM.
3.1.2.
Ozone decomposition
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results further indicated that the intermediates formed during BA degradation was capable of promoting ozone decomposition to produce HO.
3.2.3.
Effect of ozone dosage on BA degradation
Generally, ozone dose is a key parameter in ozonation process. Fig. 4 shows the degradation of BA with various doses of ozone (0.5, 1.0, 2.0, and 4.0 mg/L). The increase in ozone concentration from 0.5 mg/L to 4.0 mg/L increased the BA degradation efficiency from 60% to 95%. It is generally recognized that the ozone exposure mainly influences the degradation of the ozone-reactive pollutants (kO3 > 103 M1s1) and has little impact on that of the ozone-refractory pollutants (kO3 < 10 M1s1) (Buffle et al., 2006). However, in this study the ozone dose plays an important role in the degradation efficiency of the ozone-refractory HO probe, BA. The possible reason is that the increasing ozone dose accelerated the formation of intermediate products of BA for activating ozone to yield more HO, as elucidated below. Generally, aromatic organic compounds substituted with electron-withdrawing groups (such as eCOOH, eNO2, and eCl) are recalcitrant to direct attack of ozone. Also, we generally believe that those aromatics are refractory by ozonation at acidic pHs, where little HO would be generated due to the lack of OH. Unexpectedly, the results reported herein indicated that that ozonation may be used in effective removal of recalcitrant aromatic compounds from acidic wastewater.
3.3.
Mechanism
3.3.1.
Effect of impurities
It has been reported that OH, H2O2, electron-rich compounds (e.g, phenols, amines, alkoxylated aromatics), transition metal ions, and some metal oxides can initiate ozone to produce HO (Staehelin et al., 1982; Buffle et al., 2006; Mvula et al., 2009; Kasprzyk-Horden et al., 2003). It is noted that all solutions in this study were prepared with milli-Q water (18 MUcm) and that all the reagents are at least of ACS reagent grade. The concentration of metal ions was below the detection limit in solution by using AAS equipped with graphite furnace. Hence, the effect of such possible impurities in the reagents on ozone decomposition and HO formation should be excluded. The pre-ozonation of the milli-Q solution showed its negligible effect on the degradation efficiency of BA, also excluding the background contribution to HO formation (Fig. S2).
3.3.2.
Role of H2O2
The solution was buffered at pH 2.3 using 10 mM H3PO4, where [OH] is about 2.0 1012 M. The apparent rate constant of ozone decomposition initiated by OH is calculated to be , 1982), which is so low 3.5 1010 s1 (Staehelin and Hoigne that the role of OH in initiation of ozone could be neglected. It has been reported that H2O2 and its derivatives (e.g., HO2 ) intermediately formed during ozonation of organics influence the ozone stability in aqueous solutions (Staehelin et al., 1985; Pi et al., 2005) at neutral pH. At acidic pHs like 2.3, less than 0.3 mM H2O2 is formed (Fig. S3a), which is insufficient to induce ozone decay to produce HO, considering the low level of HO2 (pKa of ~4.8) and the slow rate constant of reaction ozone with , 1982). It H2O2 (kO3 ;H2 O2 < 0.01 M1s1) (Staehelin and Hoigne was further confirmed by the slight ozone decomposition with the addition of 10 or 20 mM H2O2 (Fig. S3b). Thus, we believe that BA degradation does not rely on the production of H2O2 under acidic conditions.
3.3.3.
Role of hydroxylated intermediates
One typical process of ozone reaction with aromatic compounds is hydroxylation. It has been proved that phenolic intermediates are responsible for promoting HO production € the et al., 2009). However, it is difficult to directly detect the (No
Fig. 3 e (a) BA degradation and (b) ozone decomposition at different initial BA concentrations. Condition: pH 2.3 ± 0.1, 25 C, [O3]0 ¼ 1.0 ± 0.05 mg/L ※ blank - 0.2 mM B 0.5 mM : 1.0 mM 7 2.0 mM = 5.0 mM 10.0 mM.
Fig. 4 e BA degradation at different ozone dosage. Condition: pH 2.3 ± 0.1, 25 C, and [BA]0 ¼ 2.0 mM.
w a t e r r e s e a r c h 7 3 ( 2 0 1 5 ) 9 e1 6
hydroxylated intermediates because they react with ozone much faster than the parent aromatics, causing their rapid consumption by ozone. For instance, the second-order rate constant for ozone reaction with 4-hydroxylbenzoic acid (4HBA) at pH 2 was 1.8 104 M1 s1 (Benitez et al., 2000). Unluckily, we cannot obtain the rate constant values for 2-HBA and 3-HBA reacting with ozone. In fact, no 2- and 3-HBAs but trace 4-HBA was detected during ozonation of 100 mM BA, as shown in Fig. S4. We assume that, if all the three HBAs are intermediately formed, the reaction of ozone with 2- and 3-HBAs should be more rapid than 4-HBA. In fact, as shown in Fig. S5, the sequence of ozone decomposition initiated by three HBAs followed 2-HBA > 3- HBA > 4-HBA. Fig. 5 shows the effect of 10 mM 2-, 3-, and 4-HBA on the degradation of BA. It was substantially accelerated within 1 min and then stopped because ozone was entirely exhausted. Apparently, all the three HBAs could increase HO exposure for BA degradation. The significant increase in HO production by the presence of HBAs was further verified by EPR study (Fig. 6), suggesting that hyroxybenzoic acids could induce ozone decomposition to produce HO at acidic pHs. Fig. S5 shows different inhibitory effect of TBA on ozone decomposition in the presence of various HBAs and phenol. TBA exhibited strong inhibition on the ozone decomposition involved by 3- and 4-HBA, moderate for 2-HBA and only mild for phenol, suggesting that a substantial fraction of ozone participated in the radical reactions, especially for 3-HBA and 4-HBA. This means that HO generation induced by hyroxybenzoic acid is similar to that by phenol (Mvula et al., 2003;
Fig. 5 e Influence of HBA on BA degradation and change of HBA during ozonation. Condition: pH 2.3 ± 0.1, 25 C, [O3]0 ¼ 1.0 ± 0.05 mg/L, [BA]0 ¼ 2 mM, [HBA]0 ¼ 10 mM, and [TBA]0 ¼ 2.5 mM.
13
Buffle et al., 2006). Upon the ozonation of phenols, HO is produced through an electron transfer from phenolate/phenol to ozone, and the HO yield of phenol was about 22% in a wide pH range from 3 to 10 (Mvula and von Sonntag, 2003). The fact that TBA decreases ozone decomposition in the presence of HBAs and phenol (Fig.S5) suggests a similar mechanism for HBAs reaction with ozone as phenol, i.e., via an electron transfer mechanism. In the hydroxylation of BA, the molar ratio for formation of 2-, 3- and 4-HBA is about 1.4:2:1 (Klein et al., 1975). Their corresponding stoichiometric ratio of reacting with ozone was about 1:1.77, 1:2.57, 1:4.16, respectively (Fig. S6). Thus, we can calculated that 0.29 mol HBA could be destroyed by per mol of ozone consumed, which is close to the stoichiometry of ozone reaction with phenol (0.33) (Mvula and von Sonntag, 2003). From the above analysis, we believe that the secondary reactions of three HBAs with ozone should be involved in the HO formation and ozone decomposition.
3.3.4.
Possible reaction pathways
Based on the above analysis and the available references on ozonation, we suggested BA ozonation at acidic pHs possibly undergoes the following reactions:
BA reacts rapidly with HO (up to ~109 M1s1) to give four isomeric HO-adduct radicals (reaction 1). Similar results were reported on hydroxylation of phenol (Klein et al., 1975; Mvula et al., 2001). The first production of HO in reaction 1 is likely to derive from instantaneous decomposition of ozone in the initial phase observed in Fig. 3b, as supported by a weak DMPO-OH signal during the decomposition of ozone alone at pH 2.3 (Fig. 6). Till now several factors have been propose for the initial production of HO, such as thermal decomposition of ozone, direct ozone-water reaction, bubble breakage in agitation process and glass container (Sotelo et al., 1987; Sehested et al., 1991; 1998; Jin et al., 2012). The transient
14
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Fig. 6 e EPR spectra of reaction between ozone with HBA. Condition: pH 2.3 ± 0.1, 25 C, [O3]0 ¼ 5.0 mg/L, [HBA]0 ¼ 10 mM, [DMPO]0 ¼ 100 mM a: DMPO; b: DMPO þ 2HBA; c: DMPO þ O3; d: DMPO þ O3 þ 2HBA; e: DMPO þ O3 þ 3HBA; f: DMPO þ O3 þ 4HBA.
HO-adduct radicals produced during hydroxylation are highly reactive with oxidants like O2, and O3 (von Sonntag, 1991, Von Sonntag, 2006). Thus, in O3/O2 aqueous solution, HO-adduct radicals would involve in the competition of O3 and O2 (reaction 2 and 4). To examine the role of O2 in BA degradation, we performed BA degradation and ozone decomposition experiments under N2-sparged (2.0 105 M O2), O2-sparged (1.0 103 M O2) and raw (2.5 104 M O2) conditions, and the results are shown in Fig. 7. At the time of 10 min, even the increase of 50-time oxygen concentration from N2-sparged to O2-sparged conditions just had mild effect on BA degradation and ozone decomposition. It suggested that HO-adduct radical reaction with ozone is dominant, even though O2 is at least in a 10-fold excess over O3 in the experimental solutions. It is reasonable because of the very low rate of HO-adduct radicals substituted with eCOOH reacting with O2 (the second-order rate constant ~106 M1s1, Fang et al., 1995). Naumov and von Sonntag (2011) suggested that ozone reacted with HO-adduct radicals by oxygen transfer based on
quantum-chemical calculations, and no HO, O3 $ or O2 $ /HO2 is generated like as during ozonation of sulfurcontaining compounds. It is not the case for ozone decomposition through a HO chain reaction during ozonation of BA. We think that ozone is unlikely to react with HO-adducts by oxygen transfer, and electron transfer seems to make sense. The HO-adduct radicals substituted by eNO2, eCN, eOCH3 and eF are oxidized to phenolic compounds by Fe3þ, Cu2þ, Cr6þ through electron-transfer (Eberhardt, 1977). Ozone has stronger oxidative ability and would be likely to react with the isomeric HO-adduct radicals by electron transfer to yield O3 $ , which is a necessary and effective precursor of HO via reaction 6 and 7. The hydroxylation species are further oxidized by O3 to produce O3 $ via reaction 3. With regard to the reactivity with oxygen, each of HOadduct radicals may in turn yield two different peroxyl radicals, namely, the 1,3- and 1,4-cyclohexadienyl-type species (Fang et al., 1995), whereby 1,3-cyclohexadienyl-type specie is only the minor component. It has been observed that rapid HO2 formation can only result from the peroxyl radicals of the cyclohexa-1,3-diene structure but not that for the cyclohexa1,4-diene structure (Pan et al., 1993). Hence, only a minor part of peroxyl radicals can eliminate HO2∙ and thereafter makes a slight contribution to HO formation via reaction 5 and 8. Thus, the pathway of electron transfer between O3 and HO-adduct radicals contributed the main source of HO, while HO2 accounts for only minor source of HO. Based on eq. (2e3), 1 mol BA would consume 4 mol of ozone in theory, i.e., 1 mol by HO-adduct radicals and 3 mol of by formed hydroxybenzoic acid via electron transfer. Fig. 2 shows that 20 mM ozone was completely decayed, whereas about 5 mM BA was degraded within 10 min. The observed stoichiometric ratio of about 1:4 for BA degradation and ozone decomposition is consistent with both equations. The intermediates produced during ozonation of BA have been shown to be capable of activating ozone to generate HO. Actually, the intermediates formed during ozonation of aromatics substituted with other EWGs such as eNO2 and eCl are also able to initiate ozone to produce HO in our study (seen in Supporting Information for review only), and those findings will be reported later. In summary, it is a common property for the aromatic compounds substituted by electronwithdrawing group such as eCN, eF, eBr, eI, eCHO, eCOOR, eNH3 þ and eSO3 . It should be noted that the proposed mechanism is not fully justified. To better understand the mechanism, further studies are still required to identify the reaction intermediate products and pathways of HO production. In addition, the roles of different electron-withdrawing groups in HO yield should be further examined.
4.
Conclusions
Based on this study, we may make the following conclusions:
Fig. 7 e Effect of oxygen on BA degradation and ozone decomposition. Condition: pH 2.3 ± 0.1, 25 C, [O3]0 ¼ 1.0 ± 0.05 mg/L, [BA]0 ¼ 10 mM.
(1) Benzoic acid was degraded efficiently by ozone via HOinduced mechanism under acid conditions, and the efficiency was significantly enhanced by the increased concentration of BA and ozone dosage.
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(2) The intermediates produced during BA degradation are involved in ozone activation to generate HO as well as ozone decomposition. (3) Under acidic conditions, the precursor of HO, i.e. O3 $ was formed mainly through the electron transfer reaction of ozone with HO-adduct radicals accounted for the major contributions, while HO2 elimination of peroxyl radicals make minor contributions (4) Hydroxylated intermediates played a positive role in the HO production, as evidenced by EPR study and products analysis.
Acknowledgment This research was supported by the Program of National Natural Science Foundation of China (21307057), the Program of Natural Science Foundation of Jiangsu Province (BK20130577), the Specialized Research Fund for the Doctoral Program of Higher Education of China (SRFDP, 20130091120014), and the Fundamental Research Funds for the Central Universities (20620140128).
Appendix A. Supplementary data Supplementary data related to this article can be found at http://dx.doi.org/10.1016/j.watres.2015.01.010.
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