Degradation of p-Acetamidophenol in Aqueous Solution by Ozonation

Jun 30, 2014 - ... in Aqueous Solution by Ozonation: Performance Optimization and Kinetics ... Journal of Environmental Management 2018 224, 376-386 ...
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Degradation of p-Acetamidophenol in Aqueous Solution by Ozonation: Performance Optimization and Kinetics Study Qizhou Dai, Jiayu Wang, Liling Chen, and Jianmeng Chen* College of Biological and Environmental Engineering, Zhejiang University of Technology, Hangzhou 310032, China ABSTRACT: The degradation of p-acetamidophenol (APAP), which is a typical component in pharmaceuticals and personal care products (PPCPs), in aqueous solution by ozonation process was studied. The effects of factors that affected the reaction rate, including the initial concentration of APAP, pH value, and O3 dosage were investigated, and the degradation products and mechanism of ozonation were also discussed. The results showed that the degradation of APAP was affected by the initial concentration of APAP, pH value, and O3 dosage. The optimized degradation condition was observed at pH 12.0. Under the same condition, increasing the pH value and ozone dosage would improve the degradation efficiency of APAP, while the degradation rate of APAP increased as the initial concentration of APAP decreased. Under the conditions of pH 2.0−12.0, an ozone dosage of 16−64 mg min−1, a pollutant concentration of 200−3000 mg L−1, and a temperature of 25 °C, the degradation followed the pseudo-first-order kinetics. Moreover, the kinetic model was proposed with the function of C = C0 exp(−11.359QO31.3475C0−0.8487[OH−]0.0343t).

1. INTRODUCTION In recent years, the pollution of pharmaceuticals and personal care products (PPCPs) has received high attention by environmental workers. With regard to pharmaceuticals, for example, there has been growing concern about the pollution occurrence of pharmaceuticals in the environment.1−5 PAcetamidophenol (APAP), which is one component of typical pharmaceuticals, is wildly applied in human medicine with annual prescriptions in China.6 APAP can be assembled from different sources, such as emissions from production sites, direct disposal of overdue APAP in households, and inadequate treatment of manufacturing waste containing APAP.5,7 Therefore, APAP is often detected in natural water, which is hazardous to humans and the environment.8,9 Therefore, it is urgent to treat this type of wastewater before it is discharged into the environment. Although conventional biological treatment is a lessexpensive method, it is hard to remove APAP, because of the toxicity of the wastewater and poor biodegradability.9−11 Ozonation has been receiving considerable attention, bceause of its strong oxidizability, its environmental friendly status, and its potential applications in environmental pollution control.12−14 Some studies focused on the degradation of APAP by ozonation and UV,15 H2O2/UV system,16 photochemical,17 Fenton oxidation,18 electro-Fenton and photoelectro-Fenton,19 and relative good results were achieved. Especially under alkaline condition, hydroxyl radical generated from ozone decomposition can react with organic compounds with typical rate constants in the range of 106−109 M−1 s−1.20 The reaction is nonselective and will not cause secondary pollution. So, ozonation is promising and utilized as a preoxidation process for the biodegradability improvement in refractory organic wastewater treatment.21−23 However, ozonation is now still mainly applied in the field of surface water treatment, and the present studies were mainly focusing on drinking water treatment with low concentration of APAP.15−19 Little study of ozonation paid attention to the pretreatment of highly © 2014 American Chemical Society

concentrated APAP pharmaceutical wastewater or APAP production workshop mother liquor before it was discharged into tradition biological treatment system. The degradation and kinetics of APAP ozonation in aqueous solutions are also limited. In this study, APAP was selected as a model pollutant and the degradation with a high concentration in aqueous solution via ozonation processes was explored. The effects of the factors that affected the reaction rate were carefully studied. Based on the intermediates detected by gas chromatography−mass spectroscopy (GC-MS), ion chromatography (IC), and highperformance liquid chromatography (HPLC), a possible degradation pathway was proposed and a kinetic reaction model of APAP degradation was deduced. This paper can provide basic data and theoretical support, which may be helpful for highly concentrated pharmaceutical wastewater treatment by ozonation.

2. EXPERIMENTAL SECTION 2.1. Reagents. P-Acetamidophenol (APAP) was selected as the model pollutant, which was purchased from Shanghai J&K Chemical Reagent (China) Co., Ltd. (purity 99%). The structure of the APAP is shown in Scheme 1. Tert-butyl alcohol (t-BuOH) was chosen as the radical scavenger and purchased from Shanghai Lingfeng Chemical Reagent (China) Co., Ltd. The pH of solutions was adjusted by sulfuric acid and Scheme 1. Structure of p-Acetamidophenol

Received: Revised: Accepted: Published: 11593

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Co., Ltd., China) and purged through a sintered porous ceramic disperser with 0.16−0.25 mm i.d. placed at the bottom of the reaction cylinder to ensure the formation of small bubbles. The flow rate was controlled by a rotameter with the saturated ozone concentration of 9.6 mg/L in the aqueous solution. The unreacted ozone in the off gas was absorbed by sodium thiosulfate solution before it was discharged into the environment. Experiments were carried out at 25 °C. The simulated wastewater containing APAP (1.5 L) was used, with ultrapure water as the solvent for the APAP solutions. After the simulated wastewater containing APAP with different pH value was added into the reactor, the reaction started when ozone addition. The reactor ran in a semibatch mode with internal circulation and samples were withdrawn at predetermined time points from the top of the reactor, and filtered through a 0.22-μm membrane filter before analysis. 2.3. Analytical Methods. Sample concentrations of APAP and other phenyl intermediates were immediately determined by HPLC (Model 1200, Agilent Technologies, USA) equipped with a C18 reverse-phase column (Eclipse XDB, Agilent Technologies, USA) and a UV detector. The UV detector wavelength was set at 254 nm. The injection volume was 5 μL. The mobile phase was methanol and water with the ratio of 15:85 (v/v). The flow rate was 0.8 mL min−1 with a column temperature of 30 °C. Other degradation intermediates were detected by gas chromatography coupled with mass spectrometry (GC/MS) (GC, Agilent 6890 system; MS, Agilent 5975 mass

sodium hydroxide. Other materials used in the test were of analytical grade. 2.2. Apparatus and Procedures. Figure 1 is our experiment systematic process diagram. The main equipment

Figure 1. Experimental equipment. Legend: 1, oxygen; 2, rotameter; 3, ozone generator; 4, gas dispersion equipment; 5, reactor; 6, bubble; 7, inner pipe of reactor; 8, outer pipe of reactor; 9, sampling connection; and 10, absorption bottle.

was a cylindrical Pyrex glass reactor. It was an internal circulation reactor (inner tube diameter = 50 mm; outer tube diameter = 80 mm; height = 650 mm; volume = 1.5 L). Ozone was generated from high-purity oxygen with an ozone generator (CFY-3, Hangzhou Rongxin Electronic Equipment

Figure 2. Effect of process variables on the degradation of p-acetamidophenol: (a) effect of pH on p-acetamidophenol degradation (T = 25 °C, C0 = 500 mg L−1, O3 dosage = 48 mg min−1); (b) effect of free radical trap on p-acetamidophenol degradation (T = 25 °C, C0 = 500 mg L−1, O3 dosage = 48 mg min−1); (c) effect of initial concentration on p-acetamidophenol degradation (T = 25 °C, pH 12.0, O3 dosage = 48 mg min−1); and (d) effect of O3 dosage (T = 25 °C, pH 12.0, C0 = 1000 mg L−1). 11594

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APAP could be removed without t-BuOH after 30 min, while after adding 10 mmol L−1 and 50 mmol L−1 of t-BuOH, the APAP removal efficiency decreased to 66% and 25% after 30 min, respectively. While at pH 2.0, the results showed that it was hardly influenced by t-BuOH. The differences between pH 12.0 and pH 2.0 showed that •OH was the main oxidizing species in the alkaline reaction, and high removal under alkaline conditions could be ascribed to •OH oxidation. 3.1.2. Effect of Initial Concentration of APAP. Figure 2c showed the effect of initial concentration of APAP on the condition of pH 12.0 and an ozone dosage of 48 mg min−1. As is shown in Figure 2c, when the initial concentration of APAP from 200 mg L−1 increased to 3000 mg L−1, after 30 min, the 100% removal of APAP decreased to 32%. Therefore, a higher initial concentration would lead to a slow degradation rate. It may due to the following two reasons. First, the number of hydroxyl radicals generated influenced the removal rate of APAP, as the experiments were performed under the same operating conditions, and the rate of generation of hydroxyl radicals was almost constant. Within the reaction time, only limited APAP was removed by ozonation. Second, the high concentration of APAP would lead to more intermediates generated, which would consume more hydroxyl radicals and thus influenced the reactions of radicals with APAP. 3.1.3. Effect of Ozone Dosage. Figure 2d showed the effect of O3 dosage on APAP degradation at 1000 mg L−1 solution and pH 12.0. The results showed that, after 30 min, the removal increased from 29% to 81% as the O3 dosage increased from 16 mg min−1 to 64 mg min−1, which indicated that the O3 dosage had a positive effect on APAP degradation. As the ozone dosage increased, ozone concentrations in the solution were increased and the mass-liquid transmission was enhanced, which resulted in an increase of the formation of the free radicals with a consequent increase in the rate of APAP degradation.26,33 The stoichiometry between O3 and APAP was 2,34 and when the O3 dosage and ozonation time were increased, the O3 utilization ratio decreased. When the ozone flow rate was higher than 48 mg min−1, the increase of the reaction rate would be very slow. 3.2. Carbon Mineralization. The degradation of APAP produces intermediates, which compete with the parent APAP for •OH and O3, etc. Consequently, destruction of the APAP should be evaluated as an overall degradation process, involving ultimately the mineralization of both the parent APAP and its intermediates. The most practical means of estimating this overall process is to monitor the TOC removal. The efficiency of APAP degradation and TOC removal in 500 mg L−1 APAP by ozonation under optimized conditions is presented in Figure 3. The results show that the efficiency of APAP degradation was nearly 100% after 60 min compared with a total mineralization of 40% after 60 min. The observation that APAP degradation was considerably faster than TOC removal was attributed to the ease of benzene ring destruction. The remaining degradation products are succinic acid, maleic acid, fumaric acid, oxalic acid, and acetic acid, etc., as seen in Table 1. The data suggest that much longer contact was necessary for appreciable conversion of TOC to CO2 and H2O. The results also show that it is not economic for the ozonation of APAP to total mineralization, and the proper way is the use of ozonation as a pretreatment process with APAP removed and wastewater biodegradability raised. After that, we can use the biological process to treat the pretreated wastewater to total mineralization.

spectrometer, USA). The gas chromatograph was equipped with a WCOT fused silica series column (30 m × 0.25 mm i.d., film thickness = 0.25 μm) and interfaced directly to the mass spectrometer. The GC column was operated at a temperature of 70 °C for 5 min and then increased at the rate of 5 °C min−1 to 280 °C and held at that temperature for 5 min. The other experimental conditions were as follows: EI impact ionization = 70 eV; helium was used as the carrier gas, at a flow rate of 1.0 mL min−1; injection temperature, 180 °C; and source temperature, 220 °C. Anion and low-molecular organic acids during degradation process were analyzed with a Dionex model ICS 2000 ion chromatograph equipped with a dual-piston (in series) pump, a Dionex IonPac AS19 analytical column (4 mm × 250 mm), an IonPac AG19 guard column (4 mm × 250 mm), and an electrical conductivity detector. The samples (2 mL) were injected manually through an injection port. The concentration of ozone in liquid phase was determined by an indigo method,24 and the ozone concentration in gas phase was determined by an iodimetric method.25 The pH value of the solutions was measured by a pH meter (LeiCi pHs-3E, Shanghai Jingke Co., Ltd., China).

3. RESULTS AND DISCUSSION 3.1. Optimization of Operational Factors. 3.1.1. Effect of Initial pH. The initial pH of wastewater is a major parameter in determining the performance of the ozonation. The effect of pH was studied on the conditions of a APAP concentration of 500 mg L−1 and ozone dosage of 48 mg min−1, as shown in Figure 2a. The results indicated that, under alkaline conditions, the degradation performance was more obvious than those of acidic and neutral conditions.26−28 At pH 12.0, 97% of the APAP could be removed after 30 min, while under the conditions of pH 2.0, 5.0, 7.0, and 9.0, the effects were 82%, 86%, 86%, and 91%, respectively. The degradation rate was increased rapidly as the the initial pH increased. This is because, under acidic conditions, APAP was removed via direct oxidation of the O3 molecule rather than hydroxyl radicals formed from ozone decomposition. Under acidic conditions (pH 4.25), the reaction was selective. So the removal by O3 was limited. As the solution became more alkaline, OH− initiated the chain reactions of ozone decomposition to hydroxyl radicals and other secondary oxidants,29,30 which greatly enhanced the degradation of APAP. So the optimal initial pH during these experiments was 12.0. The APAP would dissociate at high pH around pKa, and ozone and hydroxyl radicals would react with both the pacetamidophenol ion and the p-acetamidophenoxide ion, as the relative references mentioned,26,31,32 which would also lead the enhancement of the APAP degradation under high pH conditions. In addition, the pH values of different time intervals were analyzed during the reaction time. For example, at pH 12.0 initially, the pH would decrease during the reaction time. The results also showed that, after 30 min, the pH value would be lower than the pKa of p-acetamidophenol within 30 min, which indicated that, after 30 min, the influence of the acetamidophenoxide ion could be ignored. Radical scavenger tert-butyl alcohol (t-BuOH) was used to confirm the above mechanism in this research, because it can react with •OH rapidly with a rate constant of 5.0 × 108 M−1 s−1 and very slowly with ozone with a rate constant of 0.03 M−1 s−1.26,27 Figure 2b showed the effect of t-BuOH on APAP degradation. The results showed that, at pH 12.0, 97% of the 11595

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(D7), succinic acid (D8), acetic acid (D9), oxalic acid (D10), and formic acid (D11), were also detected by IC and HPLC. The intermediates detected in the experiments are listed in Table 1. Formic acid, acetic acid, and oxalic acid were chosen as the major intermediates to illustrate the variation in the concentration of these acids at different pH values. Figure 5 shows the variation of formic acid, acetic acid, and oxalic acid during degradation at acidic and basic pH. At pH 12.0, the concentration of formic and acetic acid was accumulated within 60 and 70 min, respectively, then it started to decrease and finally reached a stable value; the concentration of oxalic acid increased during 180 min, then reduced slowly. At pH 2.0, the concentration of formic and acetic acid was accumulated within 80 and 100 min, respectively, then started to decrease, while the concentration of oxalic acid increased during all the reaction time. The removal of intermediates reflected the removal of COD and the degree of mineralization. Under alkaline conditions, the degradation of intermediates was more obvious than those of acidic conditions, which showed that the removal of COD was more effective. Because of the alkaline condition, ozone was inclined to decompose to •OH, and the oxidation by •OH is nonselective, which enhanced the COD removal and mineralization. 3.3.2. Degradation Pathway. Based on the intermediates detected by GC/MS, IC, and HPLC (Table 1), including paminophenol (D2), p-dihydroxybenzene (D3), o-dihydroxybenzene (D4), p-benzoquinoe (D5), o-benzoquinoe (S2), maleic acid (D6), fumaric acid (D7), succinic acid (D8), acetic

Figure 3. Comparison of p-acetamidophenol degradation and TOC removal (T = 25 °C, pH 12.0, C0 = 500 mg L−1, O3 dosage = 48 mg min−1).

3.3. Intermediates and Degradation Pathway. 3.3.1. Intermediates and Variation in the Concentration of Related Acids. Many intermediates were generated during the degradation process. Figure 4 shows the results of the intermediates identified by GC-MS with theoretical and experimental m/z values under a typical condition of 500 mg/L APAP, such as p-aminophenol (D2), p-dihydroxybenzene (D3), o-dihydroxybenzene (D4), benzo-1,4-quinoe (pbenzoquinoe, D5), benzo-1,2-quinoe (o-benzoquinoe, S2), maleic acid (D6), fumaric acid (D7), and succinic acid (D8). Some intermediates, such as maleic acid (D6), fumaric acid Table 1. Intermediates Detected

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Figure 4. Representative GC/MS spectrum during APAP degradation.

Figure 5. Variation of formic acid, acetic acid, and oxalic acid during degradation.

acid (D9), oxalic acid (D10), and formic acid (D11), a possible pathway of APAP degradation was proposed, as shown in Figure 6. The hydroxyl radical played the key role during APAP degradation in ozonation. In the APAP degradation, the hydroxyl radical would attack C−N bond and be electrophilic addition on APAP at the α-position.34 For the attack of C−N bond by •OH, p-aminophenol (D2) and p-dihydroxybenzene (D3) would be produced; and when the hydroxyl radical was an electrophilic addition on APAP, S1 would be formed. While for the ozone attack on APAP, ozone would probably attack the αposition, and then S1 would be formed due to structure transformation. After that, S1 would be further oxidized to D4 by •OH or O3. Because of self-electron transfer in the molecule, D3 and D4 were transferred to D5 and S2. After the benzene ring was broken, these intermediates decomposed to succinic, maleic, and fumaric acids, which could be further decomposed to acetic and oxalic acids. These acids were gradually decomposed to formic acid and further degraded to CO2 and H2O. 3.4. Kinetics Study. As noted by Beltrán,35 ozone reactions in wastewater include gas−liquid reactions in which ozone transfers from the gas phase to the water phase, where it simultaneously reacts with other substances (pollutants) while diffusing. The Hatta number for reaction is influenced by other types of reaction conditions, including the organic concentration. Therefore, if a relatively macroscopical kinetics was established with ozone dose, organic concentration, and pH, it would be easily used in the highly concentrated organic wastewater pollution control with ozonation.

Figure 6. Probable degradation pathway of p-acetamidophenol.

Thus, the reaction kinetics of APAP by ozonation was studied and the rate constants of APAP degradation were determined under different operational conditions. Since the oxidizing ability of ozone comes from either molecular ozone or hydroxyl free radicals, the rate of APAP degradation could be formulated as follows:36 11597

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Industrial & Engineering Chemistry Research d[C] = −k 0[C][O3] − k•OH[C][•OH] dt

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(1)

where [C] is the concentration of APAP in the solution; [O3] and [•OH] are the concentrations of ozone and hydroxyl radicals, respectively; and k0 and k•OH are the kinetic rate constants, respectively. Generally, eq 1 can be simulated to the pseudo-first-order equation in eq 2, where kobs is the rate constant: d[C] = −(k 0[O3] + k•OH[•OH])[C] = −kobs[C] dt

(2)

The experimental data then was used to fit eq 2. The results of the rate constants of APAP degradation at different operational conditions fit well with the pseudo-first-order kinetics and the linear correlation coefficient (R2) is more than 0.989. Therefore, APAP degradation by ozonation is a first-order kinetic reaction with respect to APAP. Moreover, the pseudoreaction constant of kobs was 0.1219 min−1 under the conditions of pH 12.0 and 500 mg L−1 APAP. Ozonation includes both direct reaction and indirect reaction, so the liquid ozone concentration is difficult to figure out in our experiment. Therefore, we supposed that the apparent rate constant kobs is related to O3 dosage.37 In addition, from the “optimization of operational factors” in this paper, the apparent rate constant kobs is also related to the pH value, the initial concentration of APAP, and temperature. Therefore, an empirical equation (eq 3) could be used to fit the experimental data.37,38 ⎛ Ea,obs ⎞ α β kobs = A exp⎜ − ⎟Q C [OH−]γ ⎝ RT ⎠ O3 0

(3)

Equation 3 could be transformed to eq 4 by taking the logarithm: Ea,obs

ln kobs = ln A −

RT

+ α ln Q O + β ln C0 + γ 3

−

ln[OH ]

(4)

where −Ea,obs/RT is a constant, because our experiments were performed at room temperature. Then, regression analysis was performed between ln kobs and ln QO3, ln C0, ln[OH−], as is shown in Figure 7. Figure 7a was the least-squares analysis between ln kobs and ln QO3, and α = 1.3475 (from the slope of the straight line (R2 = 0.9939)). Similarly, β and γ were equal to −0.8487 and 0.0343 from Figure 7b and 7c, respectively. Moreover, their linear correlation coefficients were 0.9798 and 0.9917, respectively. Equation 4 can be converted to eq 5 based on the value of α, β, and γ: ⎛ Ea,obs ⎞ kobs = A exp⎜ − ⎟Q 1.3475C0−0.8487[OH−]0.0343 ⎝ RT ⎠ O3

Figure 7. Log−log plots: (a) ln kobs vs ln QO3, fitted curve; (b) ln kobs vs ln C0, fitted curve; and (c) ln kobs vs ln[OH−], fitted curve.

C = C0 exp( −11.359Q O 1.3475C0−0.8487[OH−]0.0343 t ) 3

To verify the accuracy of this kinetic equation, the theoretical data calculated from kinetics model were compared to the experimental data. The results showed that, under the conditions of pH 2.0−12.0, an initial APAP concentration of 1.325−19.868 mmol L−1, an O3 dosage of 16−64 mg min−1, and room temperature (25 °C), the relative deviation between experimental data and theoretical data was