Amino-Aromatic

27 May 2010 - It has great significance to investigate the degradation of aromatic contaminants as they are highly carcinogenic and nondegradable poll...
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Ind. Eng. Chem. Res. 2010, 49, 6243–6249

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Mineralization of Trace Nitro/Chloro/Methyl/Amino-Aromatic Contaminants in Wastewaters by Advanced Oxidation Processes Wenjuan Huang, Yuanhui Ji, Zhuhong Yang, Xin Feng, Chang Liu, Yinhua Zhu, and Xiaohua Lu* State Key Laboratory of Materials-Oriented Chemical Engineering, Nanjing UniVersity of Technology, Nanjing 210009, Jiangsu, P. R. China

It has great significance to investigate the degradation of aromatic contaminants as they are highly carcinogenic and nondegradable pollutants in drinking water. In this paper, the mineralization orders of the representative nitro/chloro/methyl/amino-aromatic contaminants with oxidants ( · OH, H2O2, · O-, O2,O3) in advanced oxidation processes (AOPs) are investigated based on the calculated standard molar Gibbs free energy changes of reaction (∆rG0m) and the results are consistent with those from previous experimental results, electrophilic substitution orientation rules of the Hammett equation, and predicted results with a quantitative structure-activity relationship (QSAR). In addition, the quantitative function relationship between the degradation rate (r) of the aromatic contaminants and the thermodynamic driving force (∆rG0m) is analyzed in order to investigate the degradation kinetics more rigorously. 1. Introduction Aromatic compounds are important starting materials and intermediates in the manufacturing of industrial chemicals such as dyes, pesticides, synthetic polymers, etc.1 As these widely used aromatic compounds are discharged into the atmosphere and groundwater, they constitute an important class of environmental pollutants due to their carcinogenicity, teratogenicity, and toxicity. Nowadays, much stricter drinking water standards by expanding the scope of regulated contaminants and lowering their maximum levels2 have been proposed and the treatment of drinking water is a critical issue all over the world. Traditional wastewater treatment techniques including activated carbon adsorption, chemical oxidation, biological treatment, etc. have difficulty in degrading trace organic contaminants especially for aromatic contaminants in drinking water. AOPs involving O3/ UV, H2O2/UV, Fenton, TiO2/UV, and various combinations of these processes3-6 are promising to provide an almost total degradation of trace aromatic contaminants by generating a highly reactive hydroxyl radical ( · OH). During the past decades, to provide information to optimize experimental conditions and design a large-scale photocatalytic reactor, many kinetic experiments were performed to determine the degradation reactivity of aromatic contaminants by calculating the degradation rate constants in AOPs.6-9 However, there are some difficulties to study the degradation reactivity of aromatic contaminants at the parts per billion level by experimental investigations. First, because of the different experimental conditions performed, various degradation orders were obtained for the same set of aromatic contaminants. Take Fenton oxidation of benzene derivatives for example, Edwards and Curci10 obtained the reactivity order: C6H6 > C6H5Cl > C6H5NO2 > C6H5CH3, the results of Augusti et al.9 were C6H5Cl > C6H6 > C6H5CH3 > C6H5NO2, while in the work of Davies et al.11 the order was C6H5NO2 > C6H5CH3 > C6H5Cl > C6H6. It is important to analyze the different experimental results of the degradation order. Second, for the investigated systems, the initial concentrations (parts per million level) of aromatic contaminants used in degradation experiments were much higher * To whom correspondence should be addressed. Tel.: +86-2583588063. Fax: +86-25-83588063. E-mail address: [email protected].

than those in drinking waters (low parts per billion level), which would not reveal the actual situations. For this reason, quantitative analysis of the degradation of trace aromatic contaminants still remains a great challenge.12 Kinetic models (Freundlich, Langmuir-Hinshelwood, etc.13-15), deriving the reactant concentrations during the course of the reaction from the Arrhenius rate law, are widely used to describe the degradation rate of organic species. However, the traditional degradation reaction rates are only used in the limiting case of thermodynamic equilibrium state,16 this is why the time related concentration data do not fit any rate expression exactly. Hence the traditional rate equations provide little insight into the mechanisms of the real catalytic reaction process. Actually, most catalytic reactions are in the state of thermodynamic nonequilibrium because of the strong surface effects of heterogeneous catalyst17 used in most AOPs and the effect of continuously sharing matter and energy with other systems. Therefore, the degradation rate is related not only to activation energy, but also the nonequilibrium parameter: thermodynamic driving force.18-21 In our previous work, Ji et al.7,22 investigated the mineralization abilities of 31 organic contaminants (10 chlorinated hydrocarbons, 4 brominated hydrocarbons, 11 aromatic hydrocarbons and their derivatives, 3 chloroacetic acids, and 3 chloroacetyl chlorides) with 5 oxidants ( · OH, H2O2, · O-, O2, O3) based on the ∆rG0m. The calculation results of the thermodynamic model agree with the experimental results by comparing the degradation rates. In this paper, on the basis of our previous work, the mineralization abilities of nitro/chloro/ methyl/amino-aromatic contaminants which are highly carcinogenic and nondegradable will be investigated by thermodynamic analysis. And the theoretical mineralization orders will be compared with that from electrophilic substitution orientation rule of the Hammett equation, previous experimental data, and quantitative structure-activity relationship (QSAR) predicted results. Moreover, the quantitative relation between the degradation rate (r) of the nitro-aromatic contaminants and the thermodynamic driving force (∆rG0m) will be analyzed in order to investigate the degradation kinetic equation of real irreversible heterogeneous reaction kinetics.

10.1021/ie100116c  2010 American Chemical Society Published on Web 05/27/2010

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2. Thermodynamic Analysis of Degradation Reactivity of Trace Aromatic Contaminants 2.1. Thermodynamic Driving Force. Nonequilibrium thermodynamics, as contrasted with classical equilibrium thermodynamics, are concerned with the rate of change of various thermodynamic potentials such as nonzero thermodynamic fluxes J (heat flux, mass flux, and reaction rate, etc.) and conjugate thermodynamic driving forces X.21,23-25 For the real nonequilibrium reaction processes such as degrading the trace organic contaminants in aqueous solution, the reaction rate (flux) of reaction is driven by reaction affinity A or Gibbs free energy change of reaction ∆rG (∆rG ) -A),21,24,26 which is a more available parameter. When the reaction is at equilibrium state, all thermodynamic driving forces of the chemical reaction become zero throughout the entire process.21 2.2. Molar Gibbs Free Energy Change of Mineralization Reaction. As the degradation rates of trace organic contaminants in AOPs have close relations with ∆rG0m, the values of ∆rG0m will be determined according to the calculation method used in the work of Ji et al.7,22 According to the structures of aromatic contaminants, the investigated compounds in this paper could be divided into the following four groups: nitro/chloro/ methyl/amino-aromatic compounds. Hence, ∆rG0m for the mineralization reaction of aromatic contaminants with possible oxidants ( · OH, H2O2, · O-, O2, and O3) in AOPs are calculated and listed in Table 1. The values of ∆rG0m for the reactions of the four kinds of aromatic contaminants with five possible oxidants are shown in Table 1. For the aromatic contaminants investigated in this paper, the values of ∆rG0m are all negative, which means that all the investigated aromatic contaminants are thermodynamically unstable with respect to oxidation. 3. Results and Discussion The mineralization reaction of aromatic contaminants is more likely to occur if the calculated ∆rG0m is smaller, or more negative. To determine their mineralization order of ease, ∆rG0m are sequenced from small to large values. As is shown in Table 1, for the mineralization of aromatic contaminants, it is obvious that the oxidative ability of the oxidants follows the order: · OH > H2O2 > · O- > O3 > O2, which is consistent with our previous results for the mineralization of other systems.7 According to the standard oxidation potential (volt) of the oxidants,3 we can see that · OH is a powerful oxidizing reagent with a high oxidation potential of 2.80 V, which may lead to stronger mineralization ability for the aromatic contaminants compared with those conventional oxidants such as H2O2, O2, and O3. Moreover, the oxidation potential of O3 and H2O2 are 2.07 and 1.78 V, respectively, which follows the order of O3 > H2O2. However, the calculation results of mineralization order in this paper also show that the magnitude of oxidation potential may not always represent the mineralization ability of the organic contaminants. The mineralization ability of the organic contaminants depends on not only oxidants, but also the structures and properties of the organic contaminants themselves and the mineralization products.7 The mineralization orders for nitro/ chloro/methyl/amino-aromatic contaminants are sequenced as follows by structural formula: 3.1. Comparison with Kinetic Rate Constants. There are many experimental investigations on the degradation reactivity for aromatic contaminants since they are familiar nondegradable substances, various AOPs by producing primary oxidizing species ( · OH) are nowadays available for degrading aromatic

contaminants in aqueous solutions. The rate constant (k) of reactions between aromatic contaminants and · OH is a key parameter to determine the degradation ability. In this paper, the mineralization of nitro-aromatic contaminants with · OH is selected as a representative system to compare the calculated mineralization orders with that from previous experimental data, electrophilic substitution orientation rule of the Hammett equation, and quantitative structure-activity relationship predicted results. The experimental mineralization rate constants (k) of nitro-aromatic contaminants with · OH are listed in Table 2. For comparative purposes, calculated ∆rG0m for nitro-aromatic contaminants with · OH are also listed in Table 2. It was shown that the rate constant of · OH reacting with NB was 3.9 × 109 L/mol · s, while the value of pCNB measured was only 2.6 × 109 L/mol · s. The plausible explanation29e was that · OH can substitute the benzene ring of pCNB to form many kinds of phenols such as p-chlorophenol, p-nitrophenol, 2-chloro5-nitrophenol, etc. in the process of degradation of pCNB. These phenols are more difficult to degrade which will undergo ringopening reactions to yield various kinds of carboxylic acids by further oxidation. So pCNB has a slower degradation reactivity than NB. While from experimental data in Table 2, pNT degrades faster than NB. Gao et al.30 explained that the · OH preferentially attacks the methyl group rather than the benzene ring to produce benzoic acid. The major intermediates during pNT degradation process included p-hydroxybenzoic acid, p-nitrobenzoic acid, and p-aminobenzoic acid, which can be easily oxidized. Therefore, methyl-aromatic compounds by side-chain oxidation have a higher reactivity than unsubstituted ones. Garcia Einschlag et al.29a and Dillert et al.31 also observed the same order of degradation reactivity: pNT > NB with H2O2/UV and photocatalytic process, respectively. It was explained that · OH is known to abstract H atoms from organic compounds, thus a reaction channel involving H abstraction from -CH3 might also contribute to the higher degradation rates observed for pNT, which shows that our calculated ease order for trace pNT and NB agree with the experimental results. During the photocatalytic degradation of pNA, the intermediates of p-aminophenol, p-benzoquinone, and hydroquinone,32 which can be easily oxidized in the aqueous solutions33 were observed. The structure of pNA, with a donor group (-NH2) linked via a phenyl ring to a strong acceptor group (-NO2), makes electron density on a benzene higher in the conjugated bond structures,34 that means pNA can be easily attacked by the electrophilic reagent hydroxyl radical. The degradation rate constants are sequenced from large to small. As shown in Table 2, the degradation ability for nitroaromatic contaminants with · OH follows the order pNA > pNT > NB > pCNB, which shows that the degradation order from the magnitude of ∆rG0m agrees well with the experimental results. In addition, experimental measurement of the rate constants for the reactions of · OH with nitroaromatic contaminants is relatively difficult because of the complexity of analytical

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Table 1. ∆rGm for the Mineralization Reactions of Nitro/Chloro/Methyl/Amino-Aromatic Compounds with Five Oxidants (H2O2, · OH, O2, O3, and · O-)a 0

a 0 The standard molar Gibbs free energies of formation in the gaseous phase ∆fGm (g) for aromatic contaminants are taken from ref 27; oxidants, CO2, halogen ions, and inorganic acid ions at 298.15 K are taken from ref 22; and the henry’s law constants are obtained from ref 28. The standard 0 thermodynamic properties for aqueous aromatic contaminants (∆fGm (aq)) are also calculated by the thermodynamic method used by Ji.22

methods and the high cost of experiments. Canonica and Tratnyek35 and Gramatica et al.36 developed predictive quantitative structure-activity relationship (QSAR) models to predict

degradation rate constants between various organic contaminants and · OH. The predicted degradation rate constants of various aromatic contaminants by · OH in water are shown in Table 2.

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0 Table 2. Mineralization Rate Constants and ∆rGm for Nitro-aromatic Contaminants with · OH

organic contaminants nitrobenzene (NB) p-chloronitrobenzene (pCNB) p-nitrotoluene (pNT) p-nitroaniline (pNA) a

h

29a b

Garcia Einschlag et al. Kusic et al.29h

(k/M · s) × 109 (experimental data)

different AOPs

(4.2 ( 0.4)a 4.7b 3.9c 2.3d 2.6e (8.2 ( 0.5)a 7.0f 14g

UV/H2O2 process photolysis pulse radiolysis UV/H2O2 process ozone UV/H2O2 process · OH pulse radiolysis

29b c

Pozdnyakov et al.

29c d

Buxton et al.

(k/M · s) × 109 (with · OH) (predicted data)

0 ∆rGm (kJ/mol) (with · OH) (this work)

3.9h

-1165.8724

1.3h

-1100.5728 -1201.0908 -1317.2453

h

14 29d e

Guittonneau et al.

29e f

Shen et al.

29f g

Gonzalez et al.

Van der Linde.29g

Table 3. Average Reaction Rate of the Nitroaromatic Contaminants by AOPsa organic contaminants

photocatalytic degradation rate ( · OH) r (µmol/L · min)

degradation rate r with heterogeneous Fenton (µmol/L · min)

nitrobenzene (NB) p-chloronitrobenzene (pCNB) p-nitrotuluene (pNT) p-nitroaniline (pNA)

3.96(pH)3), 4.01(pH)7), 3.31(pH)11)b 1.8753d 6.90(pH)3), 7.96(pH)11)b 7.0990e

1.037c 0.423c 1.597c

0 ∆rGm (kJ/mol)

-1165.8724 -1100.5728 -1201.0908 -1317.2453

a The average degradation rates by photocatalysis and H2O2 were obtained from the experimental data in the initial 30 min. b Dillert et al.31 [NB]o ) 100 µmol/L; [pNT]o )100 µmol/L; [TiO2] ) 1 g/L; T ) 30 °C; t ) 30 min. c Hofmann et al.33 [NB]o ) 25 ppm; [H2O2] ) 0.5 g/L; t ) 30 min; [pCNB]o ) 20 ppm; [pNA]o ) 20 ppm; [H2O2] ) 2 g/L; t ) 30 min. d Priya and Madras.41 [pCNB]o ) 20 ppm; Degussa P-25; t ) 30 min. e Gautam et al.32 [pNA]o ) 50 ppm; TiO2 suspensions; t ) 30 min.

By comparison, the degradation orders from the calculated result in this work are in accord with the experimental results and the QSAR predictive data. It verifies again the close relationships between the degradation rate and the thermodynamic driving force ∆rG0m for the investigated systems. 3.2. Electrophilic Substitution Mechanism. The degradation mechanism for oxidizing aromatic contaminants by · OH is mainly electrophilic substitution reaction.3 Classical empirical structure-activity analysis were previously performed to analyze the substituent effects. For the representative nitroaromatic contaminants (NO2-C6H4-B) listed in Table 2, the substituent groups are -CH3, -NH2, and -Cl, respectively. And they can generally be divided into two classes regarding their electrophilicity: activating groups (-NH2, -CH3) which increase the electron density of benzene ring, enhancing the reactivity of the ring, and deactivating group (-Cl) which removes electron density from the benzene ring, making electrophilic aromatic substitution reactions slower and more difficult than benzene itself.1,37 Thus pNA and pNT have higher degradation reactivity than NB while pCNB has the slowest degradation rate among the investigated nitroaromatic contaminants. Furthermore, the influence of the type of substituent groups at the aromatic ring has been quantitatively described by the Hammett equation log k/kH ) σF,38,39 where k and kH are the rate constants for the substituted reactant (NO2-C6H4-B) and unsubstituted reactant (C6H4NO2), respectively, while F and σ are the reaction and substitute constants. The parameter σ describes the degree of electron-donating or electron-accepting power by negative σ values for electron-donating groups or positive σ values for electron-withdrawing groups. The higher the absolute value, the stronger the donating/attracting properties.39 Compared with the contribution of the para substituents on the benzene ring from the value of σsNH2 -0.66, CH3 -0.17, H 0.00, Cl 0.23, NO2 0.78 in water solution40sthe degradation reactivity for nitroaromatic contaminants is pNA > pNT > NB > pCNB, which is consistent with the theoretical mineralization order of ease investigated in this paper on the basis of ∆rG0m. From Table 1, the degradation orders for chloro/ methyl/amino-aromatic contaminants also follow the electrophilic substitution orientation rule of the Hammett equation.

Figure 1. Relations between the photocatalytic degradation rate (r) and 0 ∆rGm .

0 Figure 2. Relations between the Fenton degradation rate (r) and ∆rGm .

3.3. Quantitative Relation Description of Degradation 0 Rate with ∆rGm . As shown in the above descriptions, the degradation order of the investigated aromatic contaminants based on ∆rG0m agrees with the experimental results and that from the QSAR predicted results, which shows close relations between degradation rate and ∆rG0m. Therefore, the quantitative 0 relations of the degradation rate and ∆rGm instead of that between degradation rate and concentrations will be analyzed in detail in order to investigate the degradation kinetics more rigorously. The correlation results between the experimental degradation rates (as shown in Table 3) of nitroaromatic contaminants and the thermodynamic driving force ∆rG0m are shown in Figures 1 and 2. As Table 3 has shown, the pH of the aqueous solution significantly affects the photocatalytic reaction associated with the reaction mechanism and the adsorption character-

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istics of the substrate on the TiO2 surface. According to the mechanism of heterogeneous photocatalysis, the concentration of OH- ions is critical for the generation of · OH. Thus, raising the pH of solution was proposed to enhance the number of OH- ions on the TiO2 surface and promote the photocatalytic degradation. On the other hand, the adsorption was more in basic medium than in acidic or neutral medium for the photocatalyst. Hence, the high degradation rate was obtained in basic medium in the work of Gautam et al.32 However, for the nitrobenzene, pH has a small effect on the photocatalytic degradation rate. The highest degradation rate of NB was obtained when the pH values are between 6.5 and 7.43 The explanation is that for photocatalyst TiO2, the zero point charge (pHzpc) is about 6.4. Since NB is a kind of unionizable compound, the neutral TiO2 surface seems beneficial for the adsorption of NB. It implies that NB will be adsorbed by the greatest extent on a catalyst surface under conditions in which pH ) pHzpc, and thus, the pH of maximum adsorption density for NB is around 6.4.44 Therefore, the effect of solution pH on the photocatalytic reaction is a complex subject which also depends on the nature of organic compounds. In defining the kinetic equations of nonideal systems, the rate, r, is not necessarily a linear function of the driving force. A theoretical driving force factor was introduced by Bradley45 with 1 - exp(∆G/RT), a function of the Gibbs free energy change and temperature. Sestak and Berggren46 and Satava47 extended the common rate equation (eq 3) using Bradley’s driving force. r ) k(T )f (R)[1 - exp(∆G/RT)]

(3)

where R and T denote the reacted component fraction and absolute temperature, k represents the rate constant, and f is an appropriate function applied in real kinetic calculations. The same driving force factor was applied by Pokol et al.48,49 in their attempt to construct rate equations for simple heterogeneous reactions. Though this assumption is so common that, in the majority of cases, it is not even declared explicitly.50 In the previous study, Sverjensky51 presented an empirical effective defining of the rates of crystalline solids far from equilibrium and free energy of reaction by ∆Gr ) -2.303RT log r, which is critical to understanding the rates and mechanisms of dissolution of crystalline solids in aqueous solutions and model of many geochemical, environment, and industrial processes. Hellmann and Tisserand52 reported an experimental investigation of the relation between the dissolution rate of albite feldspar and the Gibbs free energy of reaction. The data have been fitted to a rate equation: r ) k(1 - exp(∆Gr/RT)). This will help describe better the dependence of the dissolution rate on the free energy in the immediate vicinity of equilibrium. Moreover, our previous work investigated the dissolution and crystallization kinetics53-55 based on the principles of nonequilibrium thermodynamics. The investigations show that the thermodynamic driving force of a system could be described as the difference of the Gibbs free energy between a state with equilibrium.56 Hence, in the kinetic study, the “thermodynamic driving force” form contributed to a new explanation and important supplement to the conventional rate. While there are many possible forms that the relation between degradation rate and thermodynamic driving force factor can take, a simple nonlinear equation is used in this paper to describe the quantitative relation of photocatalytic degradation rate and the heterogeneous Fenton degradation rate to ∆rG0m in a real heterogeneous process:

r ) 0.0041

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exp(-0.0058∆rG0m)

(4)

r ) 0.0012 exp(-0.0056∆rG0m)

(5)

Equations 4 and 5 represent the quantitative description of the degradation kinetics of trace nitro-aromatic organic contaminants by coupling with the “thermodynamic driving force” through regression analysis. Due to the insufficient reaction rate data and that the photocatalytic degradation rates are obtained under different experimental conditions, the mathematical expressions do not satisfactorily fit the kinetic data from the regression curves in Figures 1 and 2. However, the theory of nonequilibrium thermodynamics has been proved reasonable and effective to provide the quantitative relations between degradation rates and thermodynamic driving force, which helps to give a deeper explanation and make the real irreversible heterogeneous reaction understandable further. Though the nonequilibrium thermodynamics approach is not very fruitful for making generalizations of the all the kinetic equations, further investigation about observing and measuring the degradation reaction rate (r) will be undertaken in order to more fully unravel the complexities of the r-∆Gr functional relation and the overall degradation mechanism by AOPs. 4. Conclusions In this paper, the mineralization orders of the representative nitro/chloro/methyl/amino-aromatic contaminants with five possible oxidants of · OH, H2O2, · O-, O2, and O3 in AOPs are investigated based on ∆rG0m. The theoretical analysis results are consistent with electrophilic substitution orientation rule of Hammett equation and previous experimental data and QSAR predicted results. In addition, the quantitative relation of the degradation rate of trace representative organic contaminants by coupling with the “thermodynamic driving force” (∆rG0m) is analyzed in order to investigate the degradation kinetics more rigorously. Further investigation will be undertaken in order to more fully unravel the complexities of the r-∆Gr relation and the overall degradation mechanism by AOPs. Literature Cited (1) Lauwiner, M.; Rys, P.; Wissmann, J. Reduction of aromatic nitro compounds with hydrazine hydrate in the presence of an iron oxide hydroxide catalyst. I. The reduction of monosubstituted nitrobenzenes with hydrazine hydrate in the presence of ferrihydrite. Appl. Catal., A: Gen. 1998, 172, 141–148. (2) Bhatkhande, D. S.; Kamble, S. P.; Sawant, S. B.; Pangarkar, V. G. Photocatalytic and photochemical degradation of nitrobenzene using artificial ultraviolet light. Chem. Eng. J. 2004, 102, 283–290. (3) Legrini, O.; Oliveros, E.; Braun, A. M. Photochemical Processes for Water Treatment. Chem. ReV. 1993, 93, 671–698. (4) Andreozzi, R.; Caprio, R.; Insola, V.; Marotta, A. R. Advanced oxidation processes (AOP) for water purification and recovery Marotta. Catal. Today. 1999, 53, 51–59. (5) Catalkaya, E. C.; Kargi, F. Advanced oxidation treatment of pulp mill effluent for TOC and toxicity removals. J. EnViron. Manage. 2008, 87, 396–404. (6) Pera-Titus, M.; Garcia-Molina, V.; Ban˜os, M. A. Degradation of chlorophenols by means of advanced oxidation processes: a general review. Appl. Catal., B: EnViron. 2004, 47, 219–256. (7) Ji, Y. H.; Yang, Z. H.; Ji, X. Y.; Feng, X.; Huang, W. J.; Liu, C.; Lu, X. H. Thermodynamic Analysis on the Mineralization of Trace Organic Contaminants with Oxidants in Advanced Oxidation Processes. Ind. Eng. Chem. Res. 2009, 48, 10728–10733. (8) Ollis, D. F. Contaminant degradation in water. EnViron. Sci. Technol. 1985, 19, 480–484.

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Ind. Eng. Chem. Res., Vol. 49, No. 13, 2010

(9) Augusti, R.; Dias, A. O.; Rocha, L. L.; Lago, R. M. Kinetics and Mechanism of Benzene Derivative Degradation with Fenton’s Reagent in Aqueous Medium Studied by MIMS. J. Phys. Chem. A. 1998, 102, 10723– 10727. (10) Edwards, J. O.; Curci, R. Catalytic Oxidations with Hydrogen Peroxide as Oxidant; Kluwer Academic Publishers: Dordrecht, 1992. (11) Davies, D. I.; Hey, D. H.; Summers, B. Homolytic aromatic substitution. Part XXXVI. Total and partial rate factors for the phenylation of some para-disubstituted derivatives of benzene. J. Chem. Soc. C. 1971, 268, 1–2684. (12) Li, W.; Yang, Z. H.; Feng, X.; Lu, X. H. Photocatalytic degradation of t race organics in water: sequencing and prediction of initial degradation rate. J. Chem. Ind. Eng. (China) 2007, 58, 1426–1431. (13) Fox, M. A.; Dulay, M. T. Heterogeneous photocatalysis. Chem. ReV. 1993, 93, 341–357. (14) Hsiao, C. Y.; Lee, C. L.; Ollis, D. F. Heterogeneous photocatalysis degradation of dilute-solutions of dichloromethane (CH2Cl2), chloroform (CHCl3) and carbon-tetrachloride(CCl4) with illuminated TiO2 photocatalyst. J. Catal. 1983, 82, 418–423. (15) Al-Ekabi, H.; Mayo, P. Surface photochemistry: CdS photoinduced cis-trans isomerization of olefins. J. Phys. Chem. 1985, 89, 5815–5821. (16) Madon, R. J.; Enrique, I. Catalytic reaction rates in thermodynamically non-ideal systems. J. Mol. Catal. A: Chem. 2000, 163, 189–204. (17) Parmon, V. N. Catalysis and non-equilibrium thermodynamics modern in situ studies and new theoretical approaches. Catal. Today. 1999, 51, 435–456. (18) Parmon, V.; Emeline, A. V.; Serpone, N. Glossary of terms in photocatalysis and radiocatalysis. Int. J. Photoenergy. 2002, 4, 91–131. (19) Garfinkle, M. The thermodynamic natural path in chemical reaction kinetics. Discrete Dyn. Nat. Soc. 2000, 4, 145–164. (20) Vuddagiri, S. R.; Hall, K. R.; Eubank, P. T. Dynamic modeling of reaction pathways on the Gibbs energy surface. Ind. Eng. Chem. Res. 2000, 39, 508–517. (21) Leites, I. L.; Sama, D. A.; Lior, N. The theory and practice of energy saving in the chemical industry: some methods for reducing thermodynamic irreversibility in chemical technology processes. Energy 2003, 28, 55–97. (22) Ji, Y. H.; Yang, Z. H.; Ji, X. Y.; Huang, W. J.; Feng, X.; Liu, C.; Lu, X. H. Thermodynamic study on the reactivity of trace organic contaminant with the hydroxyl radicals in waters by advanced oxidation processes. Fluid Phase Equilib. 2009, 277, 15–19. (23) Demirel, Y.; Sandler, S. I. Nonequilibrium Thermodynamics in Engineering and Science. J. Phys. Chem. B. 2004, 108, 31–43. (24) Kjelstrup, S.; Sauar, E.; Bedeaux, D.; van der Kooi, H. The Driving Force Distribution for Minimum Lost Work in Chemical Reactors Close to and Far from Equilibrium. 1. Theory. Ind. Eng. Chem. Res. 1999, 38, 3046– 3050. (25) Beard, D. A.; Qian, H. Relationship between Thermodynamic Driving Force and One-Way Fluxes in Reversible Processes. PLoS One 2007, 2, e144. (26) Rottenberg, H. Non-equilibrium thermodynamics of energy conversion in bioenergetics. Biochim. Biophys. Acta 1979, 549, 225–253. (27) Yaws, C. L. Chemical Properties Handbook: Physical, Thermodynamic, EnVironmental, Transport, Safety, and Health Related Properties for Organic and Inorganic Chemicals; McGraw-Hill: New York, 1999. (28) http://toxnet.nlm.nih.gov/cgi-bin/sis/htmlgen?HSDB. (29) (a) Garcia Einschlag, F. S.; Carlos, L.; Capparelli, A. L. Competition kinetics using the UV/H2O2 process: a structure reactivity correlation for the rate constants of hydroxyl radicals toward nitroaromatic compounds. Chemosphere 2003, 53, 1–7. (b) Pozdnyakov, I. P.; Glebov, E. M.; Plyusnin, V. F.; Grivin, V. P.; Ivanov, Y. V.; Vorobyev, D. Y.; Bazhin, N. M. Mechanism of Fe(OH) (aq)2+ Photolysis in aqueous solution. Pure Appl. Chem. 2000, 72, 2187–2197. (c) Buxton, G. V.; Greenstock, C. L.; Helman, W. P.; Ross, A. B. Critical review of rate constants for reactions of hydrated electrons, hydrogen atoms and hydroxyl radicals in aqueous solution. J. Phys. Chem. Ref. Data 1988, 17, 513–886. (d) Guittonneau, S.; Laat, J. D.; Duguet, J. P.; Bonnel, C.; Dore, M. Oxidation of p-chloronitrobenzene in dilute aqueous solution by O3 + UV and H2O2 + UV: a comparative study. Ozone: Sci. Eng. 1990, 12, 73–94. (e) Shen, J. M.; Chen, Z. L.; Xu, Z. Z. Kinetics and mechanism of degradation of p-chloronitrobenzene in water by ozonation. J. Hazard. Mater. 2008, 152, 1325–1331. (f) Gonzalez, M. C.; Martire, D. O.; Braun, A. M. Aqueous phase kinetic studies involving highly reactive species of environmental interest. Recent Res. Photochem. Photobiol. 1998, 2, 25–45. (g) Van der Linde, H. J. A pulse radiolytic study of aqueous nitroaniline solutions. Radiat. Phys. Chem. 1977, 10, 199–206. (h) Kuıˇc, H.; Rasulev, B.; Leszczynska, D.; Leszczynski, J.; Koprivanac, N.

Prediction of rate constants for radical degradation of aromatic pollutants in water matrix: A QSAR study. Chemosphere 2009, 75, 1128–1134. (30) Gao, J. Z.; Liu, Y. J.; Yang, W.; Pu, L. M.; Yu, J.; Lu, Q. F. Aqueous p-nitrotoluene oxidation induced with direct glow discharge plasma. Central Eur. J. Chem. 2005, 3, 377–386. (31) Dillert, R.; Brandt, M.; Fornefett, I.; Siebers, U.; Bahnemann, D. Photocatalytic degradation of trinitrotoluene and other nitroaromatic compounds. Chemosphere 1995, 30, 2333–2341. (32) Gautam, S.; Kamble, S. P.; Sawant, S. B.; Pangarkar, V. G. Photocatalytic degradation of 4-nitroaniline using solar and artificial UV radiation. Chem. Eng. J. 2005, 110, 129–137. (33) Hofmanna, J.; Freier, U.; Wecks, M.; Demund, A. Degradation of halogenated organic compounds in ground water by heterogeneous catalytic oxidation with hydrogen peroxide. Top. Catal. 2005, 33, 243–247. (34) Ma, H. J.; Wang, M.; Pua, C. Y.; Zhang, J.; Zhao, S. F.; Yao, S. D.; Xiong, J. Transient and steady-state photolysis of p-nitroaniline in aqueous solution. J. Hazard. Mater. 2009, 165, 867–873. (35) Canonica, S.; Tratnyek, P. G. Quantitative structure-activity relationships for oxidation reactions of organic chemicals in water. EnViron. Toxicol. Chem. 2003, 22, 1743–1754. (36) Gramatica, P.; Pilutti, P.; Papa, E. Validated QSAR prediction of OH tropospheric degradation of VOCs: splitting into training-test sets and consensus modeling. J. Chem. Inf. Comp. Sci. 2004, 44, 1794–1802. (37) Pierpoint, A. C.; Hapeman, C. J.; Torrents, A. Linear Free Energy Study of Ring-Substituted Aniline Ozonation for Developing Treatment of Aniline-Based Pesticide Wastes. J. Agric. Food Chem. 2001, 49, 3827– 3832. (38) Hansch, C.; LEO, A.; TAFT, R. W. A Survey of Hammett Substituent Constants and Resonance and Field Parameters. Chem. ReV. 199l, 91, 165–195. (39) Krygowski, T. M.; Zachara-Horeglad, J. E. Resonance-assisted hydrogen bonding in terms of substituent effect. Tetrahedron 2009, 65, 2010–2014. (40) Takahata, Y.; Chong, D. P. Estimation of Hammett Sigma Constants of Substituted Benzenes Through Accurate Density-Functional Calculation of Core-Electron Binding Energy Shifts. Int. J. Quantum Chem. 2005, 103, 509–515. (41) Priya, M. H.; Madras, G. Photocatalytic degradation of nitrobenzenes with combustion synthesized nano-TiO2. J. Photochem. Photobiol., A: Chem. 2006, 178, 1–7. (42) Huang, H. H.; Tseng, D. H.; Juang, L. C. Heterogeneous photocatalytic degradation of monochlorobenzene in water. J. Hazard. Mater. 2008, 156, 186–193. (43) Bhatkhande, D. S.; Pangarkar, V. G.; Beenackers, A. A. C. M. Photocatalytic degradation of nitrobenzene sing titanium dioxide and concentrated solar radiation: chemical effects and scaleup. Water Res. 2003, 37 (6), 1223–1230. (44) Yang, Y. X.; Ma, J.; Qin, Q. D.; Zhai, X. D. Degradation of nitrobenzene by nano-TiO2 catalyzed ozonation. J. Mol. Catal. A: Chem. 2007, 267 (1-2), 41–48. (45) Bradley, R. S. The energetics and statistical mechanics of the kinetics of solid fwdarw. solid reactions. J. Phys. Chem. 1956, 60, 1347– 1354. (46) Sesta´k, J.; Berggren, G. Study of the kinetics of the mechanism of solid-state reactions at increasing temperatures. Thermochim. Acta 1971, 3, 1–12. (47) (a) Satava, V. Mechanism of solid state reactions from thermal analysis measurements. In Proceedings of the 3rd International Conference on Thermal Analysis; Wiedemann, H. G., Ed.; Birkhaeuser: Basel, 1972, Vol. 2, p 373. (b) Satava, V. Fundamental principles of kinetic data evaluation from thermal analysis curves. J. Therm. Anal. 1973, 5, 217. (48) Pokol, G.; Ga´l, S.; Sztatisz, J.; Domonkos, L.; Pungor, E. Proceedings of the 8th International Symposium on the ReactiVity of Solids; Wood, J., Lindqvist, O., Helgesson, C., Vannerberg, N. G., Eds.; Plenum Press: New York, 1977. (49) Pokol, G.; Ga´l, S.; Pungor, E. Description of the rate of heterogeneous chemical reactions. Thermochim. Acta 1979, 33, 259–265. (50) Pokol, G. The thermodynamic driving force in the kinetic evaluation of thermoanalytical curves. J. Therm. Anal. Calorim. 2000, 60, 879–886. (51) Sverjensky, D. A. Linear free energy relations for predicting dissolution rates of solids. Nature 1992, 358, 310–313. (52) Hellmann, R.; Tisserand, D. Dissolution kinetics as a function of the Gibbs free energy of reaction: An experimental study based on albite feldspar. Geochim. Cosmochim. Acta 2006, 70, 364–383.

Ind. Eng. Chem. Res., Vol. 49, No. 13, 2010 (53) Cheng, F. Q.; Bai, Y.; Liu, C.; Lu, X. H.; Dong, C. Thermodynamic analysis of temperature dependence of the crystal growth rate of potassium sulfate. Ind. Eng. Chem. Res. 2006, 45, 6266–6271. (54) Ji, Y. H.; Ji, X. Y.; Liu, C.; Feng, X.; Lu, X. H. Modelling of mass transfer coupling with crystallization kinetics in microscale. Chem. Eng. Sci. 2010, 65, 2649–2655. (55) Ji, Y. H.; Qian, H. L.; Liu, C.; Ji, X. Y.; Lu, X. H. Modeling of specific structure crystallization coupling with dissolution. Front. Chem. Eng. China. 2010, 4, 52–56.

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(56) Ji, Y. H.; Lu, X. H.; Yang, Z. H.; Feng, X. Thermodynamic analysis on theoretical energy consumption for organic contaminants removal by physical methods. Sci. China Chem. 2010, 53 (3):), 671–676.

ReceiVed for reView January 18, 2010 ReVised manuscript receiVed March 29, 2010 Accepted April 28, 2010 IE100116C