Degradation of bisphenol A by ozonation and determination of degradation intermediates by gas chromatography–mass spectrometry and liquid chromatography–mass spectrometry

Chemical Engineering Journal 220 (2013) 6–14

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Degradation of bisphenol A by ozonation and determination of degradation intermediates by gas chromatography–mass spectrometry and liquid chromatography–mass spectrometry Erdal Kusvuran ⇑, Deniz Yildirim University of Cukurova, Faculty of Sciences and Letters, Department of Chemistry, 01330 Balcali, Adana, Turkey

h i g h l i g h t s

g r a p h i c a l a b s t r a c t

" The stoichiometric ratio between

ozone and BPA was calculated to be 10.30 at pH 3.0. " Mineralization of the BPA solution is not as successful as the oxidative degradation. " Ten oxidation products are identified during BPA ozonation. " Henry’s constant, kH, increased 5.5% depending on growth of ionic strength of BPA solution.

a r t i c l e

i n f o

Article history: Received 4 August 2012 Received in revised form 12 December 2012 Accepted 19 January 2013 Available online 29 January 2013 Keywords: Bisphenol A Ozonation LC–MS/MS GC–MS

a b s t r a c t In this study, bisphenol A (BPA), an endocrine disrupter, was removed by ozonation process. During ozonolysis, degradation kinetics and degradation intermediates of BPA were determined and degradation stoichiometry was also calculated. The degradation of BPA was found to be optimal at pH 3.0 and BPA solution with 0.509 mM concentration was completely degraded after 25 min ozonation time. The stoichiometric ratio between ozone and BPA were calculated to be 10.30. The pseudo-first order degradation rate constant, kObs, decreased in the range of 19.3–13.3 s1 when the initial concentration of BPA was raised from 0.051 to 0.509 mM. In addition, the second order rate constant, kapp(BPA), was also calculated in the range of 2.18  104–3.56  104 M1 s1. Henry’s constant, kH, increased as 5.5% depending on growth of ionic strength of BPA solution during ozonation. As a result of the increase of kH, dissolved ozone reduced throughout the ozonation. Ten different intermediates occurred during the ozonation of BPA and were identified via GC–MS and LC–MS/MS. Malonic and oxalic acids were observed among 0 the intermediates in the first 5 min of ozonation and taken as markers of mineralization. kTOC value 1 1 for BPA mineralization and the mineralization of ozonation was was calculated as 2.11 M min achieved about 30% at the end of 25 min ozonation. Ó 2013 Elsevier B.V. All rights reserved.

1. Introduction 2,2-Bis(4-hydroxyphenyl)propane (bisphenol A, BPA) is a dimer of two para-hydroxyphenyls bonded through a methylene bridge and indispensable monomer for the production of various polymeric materials such as polycarbonate, epoxy resins, polyacrylates ⇑ Corresponding author. Tel.: +90 322 551 20 57; fax: +90 322 551 22 55. E-mail address: [email protected] (E. Kusvuran). 1385-8947/$ - see front matter Ó 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.cej.2013.01.064

since the polymerization of the para-hydroxyphenyl rings of BPA can be easily controlled. Hence, the production of BPA has been found increasing market demand. For example, 7.6  108 and 5.6  107 kg of BPA are produced in the United States and China, respectively per year [1]. Notwithstanding its high commercial value, it has been shown that BPA is an endocrine disruptor even at a low concentration (1.0–10.0 mg L1) and can cause to detrimental health problems on people such as carcinogenesis [2–5]. Therefore, governments limited legally total BPA intake around 50 lg per

E. Kusvuran, D. Yildirim / Chemical Engineering Journal 220 (2013) 6–14

kilogram of body weight per day [6]. It was reported that BPA was detected in municipal wastewater on account of untreated industrial wastewater and demonstrated that a low amount of BPA was also very harmful for aquatic ecosystem [7]. A number of methodologies based on physical and chemical treatment processes are currently continued to apply the removal of BPA [8–14]. Advanced oxidation processes (AOPs) have been regarded as a new technology for the decomposition of several organic matters. These processes include ozonolysis [15–19], Fenton [20], photocatalysis [21–23] and wet-air oxidation methods [24]. Ozonation process is a faster and cheaper process than photocatalysis and wet-air oxidation, respectively and besides, no extra chemicals are required like Fenton process. When considering the overall factors, ozonation has provided the highly affirmative results among the AOPs [25]. Recently, the degradation of BPA using ozonation has been described by several researchers [26–30]. These studies presented significant information about BPA-ozone stoichiometry and degradation kinetic of BPA, however, none of them took into account of Henry’s constant which is an important parameter affecting onto solubility of ozone. Furthermore, a few studies were conducted for analyzing the degradation intermediates of BPA using liquid chromatography–mass spectrometry [31,32]. However, complete chromatographic identification of degradation intermediates is often difficult and requires more chromatographic techniques to identify degradation intermediates completely due to the probable existence of overlapping or embedded peaks [33]. The combination of gas chromatography–mass spectrometry and liquid chromatography–mass spectrometry is a powerful technique for evidence of degradation intermediates and provides an effective way for the quantification and identification of intermediates. Therefore, this study aimed to scrutinize degradation and mineralization processes of BPA by ozone and its degradation intermediates by gas chromatography–electrospray ionization mass spectrometry (GC– EI–MS) and liquid chromatography–electrospray ionization tandem mass spectrometry (LC–ESI–MS/MS).

7

Separation Products, Spectra System UV 2000) and a C18 intersil ODS-3 column (4.6  150 mm) at 278 nm. The mobile phase was methanol/water mixture (60/40, v/v) at a flow rate of 1.0 mL min1. HPLC detection limit of BPA was calculated as 5.93 (Signal/Noise) for 0.5  102 mM BPA concentration. These experiments were replicated three times. Ozonation experiments were carried out with 0.5 L BPA solution in a 1.0 L cylindrical glass reactor. BPA solutions were fed with ozone gas by diffusing from the bottom of the reactor. Ozone traps containing 0.5 L of KI solution (20 g L1) were connected at the outlet of this reactor to capture the unreacted ozone. Excess ozone in this traps were detected using an iodimetric method according to a previously reported procedure [34,35]. In order to detect amount of ozone produced during ozonation at a constant O2 gas velocity, 0.5 L of KI solution (pH 3.0) was measured into the reactor. The content of trap and reactor was titrated with Na2S2O3 (0.010 mol L1). The consumed ozone by BPA was determined by using the equation below:

O3Consumed ¼ O3Total  ðO3Trap þ O3Residual þ O3Headspace Þ

ð1Þ

where Consumed O3 is the amount of ozone interacted with BPA molecules, Total O3 is the ozone generated during an experiment and Trap O3 is the ozone captured by the trap. The Residual O3 and Headspace O3 is the residual dissolved ozone and the amount of ozone collected above the reaction mixture. O3Residual and O3Headspace are very small when compared with O3Trap. Therefore, they can be neglected and the final equation can be written as shown below:

O3Consumed ¼ O3Total  O3Trap

ð2Þ

One milliliter samples were taken at 0, 1, 2, 3, 5, 7, 10, 15, 20 and 25 min for cumulative ozonation times and analyzed by HPLC during ozonation of BPA. To determine the total organic carbon content (TOC) of the BPA solutions, 10 mL of samples withdrawn from the reaction mixture were analyzed by Tekmar Dohrmann Apollo 9000 instrument for all BPA initial concentrations after 25 min ozonation time.

2. Materials and methods 2.1. Materials Bisphenol A (99%, BPA) was supplied from Riedel-de Haen AG (Germany). Potassium iodide (KI), sodium thiosulfate (Na2S2O3) and 1-(Trimethylsilyl) imidazole were purchased from Merck KgaA (Darmstadt, Germany). Ozonation experiments were carried out by using an Ozo-1VTT model ozone generator (Ozomax, Canada) combined with pure oxygen gas (99.9%). The flow rate of oxygen gas was adjusted with a flow meter settled on the tube. The ozone gas flow rates and concentration of ozone gas used during all experiment were 720 mL min1 and 0.239 mmol min1. 2.2. Ozonation experiments The BPA solutions were prepared in the range of 0.051–0.509 mM concentration using distilled water. The most concentrated BPA solution was obtained by placing excess BPA into distilled water (5.0 L) at pH 10.0 and the solution was stirred with a magnetic stirrer for 1 h. Using this method, 2.0 g of BPA was successfully dissolved in the water. The BPA solution was stirred on a magnetic stirrer for an additional hour again after 1.0 M H3PO4 solution (0.5 L) was added. The desired pHs of BPA solutions were adjusted by using concentrated NaOH and H2SO4 solution. These solutions were then filtered through 0.1 lm filter paper and stored in a dark bottle at ambient temperature (20 °C) until used for ozonation reactions. The amount of BPA in this solution was detected by using a liquid chromatography instrument equipped with UV detector (HPLC- UV, Thermo

2.3. Gas chromatography–mass spectrometry (GC–EI–MS), liquid chromatography (LC–UV) and liquid chromatography–mass spectrometry (LC–ESI–MS/MS) analyses for the degradation products of BPA The intermediates occurred during ozonolysis of BPA were determined by GC–EI–MS (Shimadzu, 2010) combined with a RESTEK Rtx-5MS (0.25 lm, 60 m  0.25 mm, 5% diphenyl-95% dimethyl polysiloxane) column and an electron impact (EI) detector (70 eV). The temperature program was an initial temperature of 50 °C, held for 9 min, then increased at 5 °C min1 to 240 °C and held for 12 min. The temperatures of the ion source, injection port and interface were 230, 260 and 250 °C, respectively. To determine the degradation products of BPA, the ozone was fed into the reaction mixture by diffusion after 0.5 L of BPA solution (0.509 mM in pH 3.0 phosphate buffer) was measured into the reactor. After the BPA solution was ozonized for 5 min, 250 mL of this solution was withdrawn into a separation funnel. Subsequently, 2 mL saturated NaCl solution was pipetted and 75 mL diethyl ether was measured into the funnel before the solution was vigorously shaken. Addition of diethyl ether was repeated two more times and organic phases were combined in Erlenmeyer flask. The contents of Erlenmeyer flask were concentrated to 5 mL after drying with solid anhydrous Na2SO4. Fifty microliters of ntrimethylsilylimidazole (Supelco) was used for silylation of the product of BPA ozonized and analyzed by GC–EI–MS. The same procedure was applied for characterizing the intermediates of the ozonation reaction at the other times (10, 15, 20 and 25 min).

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The degradation products of BPA were also analyzed using LC– UV and LC–ESI–MS/MS. LC–UV analyses performed on an ODS-3 C18 column (4.6  150 mm) at 278 nm. The mobile phase was methanol/water mixture (60/40, v/v) at a flow rate of 1.0 mL min1. LC–MS analyses achieved with a triple quadrupole instrument using electro spray ionization mode (Thermo Scientific, TSQ Quantum Access Hyper Quads Model). The column was a Hypersil gold C18 column (2.1  50 mm) packed with 1 lm spherical particles. The mobile phase was methanol/water mixture (60/ 40, v/v) at a flow rate of 0.250 mL min1.

During this process, the small amount of ozone bubbles dissolves in solution while the big ones leave from reaction mixture. As a result, in basic pH, dissolved ozone produces no much more OH radicals compare with in molecular ozone in acidic pH. It was reported that ozone molecules consumed by decomposition reactions in basic solution [38,39]. Moreover, because OH radicals are non-selective, they can be consumed with other OH radical to yield hydrogen peroxide and further self decomposition reactions may be occurred. The degradation pH of BPA was therefore determined to be 3.0 and all further experiments were conducted at this pH.

3. Results and discussion

3.2. Ozone-BPA stoichiometry

3.1. The effect of pH on ozonolytic degradation

In this section, the stoichiometric relationship between ozone and BPA was examined for degradation of each initial BPA concentration and the obtained results were demonstrated in Table 1 and Fig. 2. When the decomposition percentage was achieved over 99, the ozone consumption was calculated via Eq. (1). The longest ozonation time was detected as 25 min from Fig. 1. After this time, ozone consumed by 0.509 mM initial BPA concentration was calculated as 54% of total ozone (5.982 mmol) produced. The shortest ozonation time was observed as 10 min for 0.051 mM BPA initial concentration. The ozone consumed by BPA at this concentration was calculated as 14.9% of the total ozone (2.393 mmol). The stoichiometric ratio between ozone-BPA was, therefore, calculated to be 10.30 from Fig. 2. Garoma and Matsumoto [30] reported that the stoichiometric ratio of ozone to BPA varied in the range of 7.0–9.5 in the ozonation time range of 20–30 min. Therefore, the ozone-BPA reaction can be written as:

The ozonation experiments were carried out with 0.509 mM BPA solution in phosphate buffer at pH values of 3.0, 7.0, and 10.0 for 25 min ozonation time. Degradation of BPA at the neutral and alkali pH conditions was determined as 87.0% and 94.0%, respectively, but 99.5% decomposition was obtained at pH 3.0 (Fig. 1). The effect of pH on the reaction between ozone and substrate is well known. The reaction of organics with ozone is directed in two ways by pH. One of them is direct attack or electrophilic attack (at acidic pH) and the other is mainly free radical attack (at basic pH) [36]. Moreover, the amount of ozone reacted with organics is limited by dissolved ozone in the reaction mixture. If the reaction between dissolved ozone and organics carried out in an instant of time zone is assumed a cycle, the number of cycle will increase depending on initial organic concentration. In a cycle, dissolving ozone in reaction mixture and reaching equilibrium will need time [37]. The gas form of ozone firstly inputs bottom in reaction medium, then it bubbles from bottom to up in the reaction mixture.

0.6 0.5

pH=3.0 pH=7.0 pH=10.0

[BPA] (mM)

0.4 0.3 0.2 0.1 0.0

0

5

10

15

20

25

t (min) Fig. 1. Ozonolytic degradation of BPA at various pH values ([BPA]0 = 0.509 mM).

BPA þ 10:3O3 ! OP where OP is the oxidation products. 3.3. The degradation products of BPA during ozonation process The formed intermediates during ozonation of BPA were silylated to derivatize the non-volatile organics and then identified via GC–MS. The results are demonstrated in Fig. 3 and Table 2. Fig. 3 shows the total ion chromatogram (TIC) belonging to BPA (0.509 mM) treated with ozone for 5 min. In this chromatogram, nine peaks were identified. The decrease of BPA concentration in solution was followed by LC–UV at 278 nm (Fig. 4). Seven peaks belonging to intermediates of the degradation of BPA were detected in the HPLC chromatogram after 5 min ozonation time for 0.509 mM initial concentration. However, in the ozonation time of 25 min, the three peaks with retention times (tR) 6.32, 4.49 and 3.92 min disappeared and a new peak was observed at tR 2.66 min. In addition, the peak for BPA (tR 13.05 min) was disappeared completely after this time (Fig. 5). The identification of two more degradation products (PN10 and PN11) which was not observed by GC–EI–MS was achieved with LC–ESI–MS/MS (Table 2). The possible higher compounds, such as primary oxidation products (PN10 and PN11) mentioned in Fig. 6, were not observed by GC–EI–MS analysis since they were

Table 1 The amount of consumed ozone by BPA molecules during ozonolysis experiments for different initial BPA concentrations, titrated volume of KI trap contents and titrant volume.

a

Initial [BPA] (mmol)

Ozonation time (min)

Volume (mL) (trap content)

Titranta vol. 0.010 M Na2S2O3

Excess O3 in trap (mmol)

Consumed O3 by BPA (mmol)

0.509 0.381 0.254 0.127 0.051

25 25 20 15 10

25 32 37 42 43

551.40 726.56 635.14 547.62 406.98

2.757 3.633 3.176 2.738 2.035

3.225 2.349 1.610 0.851 0.358

The trap contents titrated by 0.010 M Na2S2O3 were normalized to a reaction volume of 0.5 L.

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5

nO =10.30n[BPA]

4

2

3

r =0.999 3

3

nO (mmol)

degradation times of BPA were inversely proportional to the initial concentration of BPA. During the oxidative degradation of BPA for the 0.051 mM initial concentration, the peak of BPA in the HPLC chromatogram disappeared at the end of the 7 min ozonation whereas the degradations of BPA for 0.127, 0.254, 0.381 and 0.509 mM initial concentrations were >99.9%, 90.2%, 84.5%, 77.4% and 71.9%, respectively. As mentioned before, the ozone first attacks BPA for addition and cleavage of the double bonds to produce various intermediates then also began to react with these degradation intermediates containing double bonds in their chemical structures (Table 2). Therefore, the reaction kinetic between BPA and ozone can be written as given in the following equation:

2

1

0 0.00

0.05

0.10

0.15

0.20

0.25

0.30

0.35

n[BPA] (mmol) Fig. 2. Stoichiometric relationship between ozone and BPA.



d½BPAT ¼ ½O3 ðk1 ½BPA þ k2 ½BPA  þ k3 ½BPA2  dt

ð3Þ

where [BPA]T is the total BPA concentration including neutral and anionic species of BPA and [BPA]T can be written as:

½BPAT ¼ ½BPA þ ½BPA  þ ½BPA2  either non-volatile or thermally unstable. As a result of combining the evaluation of GC–EI–MS and LC–ESI–MS/MS results, a degradation mechanism was proposed for ozone-BPA reaction and presented in Fig. 7. The reaction of molecular ozone with carbon–carbon double bonds yields the ozonides formed via addition of ozone in double bond. Ozonides are well known unstable intermediates and degraded rapidly to carbonyl groups [40]. During the ozonation of BPA, malonic acid and oxalic acid was observed in the first 5 min. The presence of these acids showed that the mineralization of the intermediates began simultaneously with the start of ozonation of BPA. Unlike oxalic acid, malonic acid then disappeared on TIC after 5 min ozonation. The malonic acid was assumed to have been consumed by other intermediates since its degradation by ozone is not favored over degradation of BPA or intermediates containing carbon–carbon double bonds. Additionally, PN4 disappeared after 5 min of ozonation treatment, the same as for malonic acid. The areas of PN3, 6, 7, and 9 were observed to decrease with increases in ozonation time, although the area of PN5 remained at the same level. The area of PN8 increased until up to 15 min, then decreased between 15 min and 20 min and increased again at 25 min. 3.4. Kinetics of BPA ozonation The effect of initial BPA concentrations on the ozonolytic degradation of BPA was investigated in the BPA concentration range from 0.051 to 0.509 mM at pH 3.0. Fig. 8 shows the degradation of BPA at various initial concentrations. According to Fig. 8, the

ð4Þ 

2

At pH 3.0, the concentrations of [BPA ] and [BPA ] species can be omitted (Ka1 2.5  1010 and Ka2 6.3  1011 for BPA) and [BPA]T = [BPA], hence, Eq. (3) can be simply written as:



d½BPA ¼ k½BPA½O3  dt

ð5Þ

where k is the rate constant of BPA-ozone reaction and [O3] is the ozone concentration. On the other hand, the O3 concentration consumed in the liquid phase is determined using Eq. (2).

½O3Consumed  ¼ ½O3Total   ½O3Trap  Since only dissolved ozone in reaction mixture reacted with BPA, the effective ozone concentration in a liquid phase depended on partial pressure and the Henry constant. Accordingly, dissolved ozone could be calculated as shown in the following equation:

½O3 L ¼

P kH

ð6Þ

where [O3]L is concentration of ozone in liquid phase, P is the partial pressure of ozone and kH is the Henry constant of ozone. For ozone, the Henry’s constant varies depending on temperature, ionic strength and pH in a solution [41–43]. When the temperature and pH held constant for a reaction media, its ionic strength becomes a dominant factor which affects on ozone solubility. During ozonation of diluted solution, ionic strength occurred with degradation of substrate may be ignored if substrate concentration is prepared level of milimolar [44]. Finally, Eq. (7) is obtained when Eq. (6) is written into a second order kinetic equation.

Fig. 3. The GC–EI–MS total ion chromatogram for 0.509 mM BPA concentration treated with ozone for 5 min.

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Table 2 Organic intermediates identified by GC–EI–MS and LC–EIS–MS/MS. Peak no.

Name

1

Oxalic acid

Open structure

O OH

HO O 2

Malonic acid

3

1,2-Dihydroxy benzene

HO

OH O

O

OH OH

4

2-(p-Hydroxyphenyl)-2-propanol

5

4-Hydroxy-benzoic acid

OH

HO

O OH HO

6

1,4-Dihydroxy benzene

7

Trihydroxybenzene

OH HO OH HO

OH 8

O

4-(p-Hydroxy phenyl),4,4-dimethyl,1-buten-3-on

HO 9

BPA

HO



10

4-(2-(4-Hydroxy phenyl)propan-2yl)cyclohexa-3,5-diene-1,2-dione

11

(2E,4Z)-3-(2-(4-Hydroxyphenyl)propan-2-yl)hexa-2,4-dienedioic acid

  d½BPA P ¼ k½BPA dt kH

ð7Þ

When P and kH are assumed as constant under the experimental conditions, the final kinetic equation can then be arranged as shown in Eq. (7).



d½BPA ¼ kObs ½BPA dt

kObs ¼ k

P kH

ð8Þ

ð9Þ

OH

where kObs is the observed pseudo-first order rate constant, and can be calculated from the slope of the plot of ln([BPA]/[BPA]0) versus time. Plots of the linear form of the pseudo-first order kinetic model are given in Fig. 8. For initial BPA concentrations of 0.051– 0.509 mM, the first order rate constants were in the range from 13.3 to 19.3 s1. These results revealed that the first-order model gave a good fit and the coefficient of determination was determined to be between 0.991 and 0.999. As confirmation, Garoma et al. [30] also reported that they observed linearism between ln([BPA]/ [BPA]0) and time. In addition, it was noted that kobs was inversely proportion to the initial concentrations of BPA. Furthermore, when the initial concentration of BPA increased from 0.051 to 0.509 mM,

E. Kusvuran, D. Yildirim / Chemical Engineering Journal 220 (2013) 6–14

11

Fig. 4. HPLC chromatogram of the solution of BPA for first 8 min ozonation ([BPA]0 = 0.509 mM).

PN7 PN6 PN2 PN3 PN4

Concentration (mM)

0.5

kobs decreased from 19.3 to 13.3 s1. Partial pressure of ozone in reaction media may be assumed as a constant because ozone is fed continuously and k is already a constant. In this case, kH may decrease depending on increase of ionic strength. Therefore, an inversely proportional relationship was observed between first-order rate constant and BPA concentration in this study. On the other hand, BPA-ozone degradation rate can be calculated using the following equation:

PN8 PN1 PN5 PN9

0.4 0.3

½BPAT;0 ½PhenolT;0 kappðBPAÞ ln ¼ ½BPA kappðPhenolÞ ½Phenol

0.2

ln

0.1

As simply indicated in Eq. (4) ([BPA]T = [BPA]), kapp(BPA) value in Eq. (10) can be rewritten as:

0.0

0

5

10

15

20

25

Time (min)

ð10Þ

k ¼ kappðBPAÞ Finally, when k value is replaced in Eq. (9) and the equation is rearranged, Eq. (11) is obtained,

Fig. 5. The concentration variation of BPA and intermediates during ozonation.

Fig. 6. The LC–ESI–MS/MS total ion chromatogram for 0.509 mM BPA concentration at the end of 5 min ozonation to identify PN10 and PN11.

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Fig. 7. Proposed degradation mechanism for BPA.

kObs kappðBPAÞ

¼

P kH

ð11Þ

kapp(BPA) represents the rate constant of the reaction between molecular ozone and molecular BPA. kapp(BPA) value was calculated in the range of 2.18  104–3.56  104 M1 s1 via Fig. 9. Lee and Gunten [16] reported that the second order rate constant of BPA was 2.4  104 M1 s1 by using quantitative structure-activity relationship model. The kObs/kapp(BPA) ratios for each BPA concentration were observed as 5.42  104, 5.39  104, 5.33  104, 5.26  104 and 5.18  104 M, respectively. When P/kH ratios were plotted versus BPA concentrations (Fig. 10), a straight line with a slope value of 0.05 and an intercept value of 5.46  104 M was observed. The

value of 5.46  104 M represented the P/kH ratio when the concentration of [BPA] = 0 mM. This value also belonged to ionic strength of the buffer solution in the absence of BPA and indicated that there was no ionic strength to be derived from BPA. When [BPA] = 0 and [BPA] = 0.509 mM, the corresponding kH (atm M1) values were calculated as 1.83  103 P and 1.93  103 P indicating that kH value increased 5.5% with concentration increase. This finding can be explained due to increasing of the ionic strength of solution during BPA ozonation. Andreozzi et al. [41] investigated the relationship between Henry’s constant and ionic strength and they reported that Henry’s constant was a proportional to ionic strength. Depending on the increment of Henry constant, the solubility of ozone reduced and this caused the change of first order and second order rate constants.

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0,5 0.5

[BPA]0 (mM)

0,4 0.4

-0.30

4

Y = 2.11X-0.63

3

ln (CTOC25/CTOC0)

-ln ([BPA] / [BPA]0)

5

2 1 0

0.3 0,3

0

5

10

15

20

25

t (min)

[BPA]0=0.509 mM [BPA]0=0.381 mM [BPA]0=0.254 mM [BPA]0=0.127 mM [BPA]0=0.051 mM

0,2 0.2 0,1 0.1

2

r =0.998 -0.35

-0.40

-0.45 0.10

0

5

10

15

20

0.12

25

0.14

0.16

25.CO3 (M.min)

t (min)

0

Fig. 8. Degradation of BPA in solutions of various initial concentrations by ozonation (pH 3.0) and the linear transform ln([BPA]/[BPA]0) versus time.

Fig. 11. Plot of lnðC TOC 25 =C TOC0 Þ versus tCo3 to determine kTOC value (t = 25 min).

3.5. Mineralization of BPA

7 6

ln ( [BPA]T,0/ [BPA] )

The mineralization of BPA was measured using TOC analyzer. After 25 min ozonation time, TOC value reduced about 30% while initial BPA concentration was completely depleted at the end of 25 min. This finding shows that the reaction medium contains saturated organic compounds such as carboxylic acids. The mineralization of BPA was also estimated from the equation according to Rivas et al. [29].

kapp(BPA) = 1.67.kapp(Phenol) kapp(BPA) = 1.93.kapp(Phenol) kapp(BPA) = 2.23.kapp(Phenol) kapp(BPA) = 2.59.kapp(Phenol) kapp(BPA) = 2.74.kapp(Phenol)

8

5 4



dCTOC ¼ kTOC 1 CTOC aCO3 ; dt



dCTOC 0 ¼ kTOC CO3 dt CTOC

3 2

COH CO3

ð12Þ

ð13Þ

When this equation was integrated, Eq. (14) was obtained.

1



0 0.0

0.5

1.0

1.5

2.0

2.5

3.0

ln ( [Phenol]T,0/ [Phenol] ) Fig. 9. Linear transform of ln([BPA]T,0/[BPA]) versus ln([Phenol]T,0/[Phenol]) to determine kapp(BPA)/kapp(Phenol) ratio at different initial [BPA] concentration levels.

6 5 4

p/kH=-0.05 [BPA)]+5.46

3

r =0.99

2

2 1 0

Z

CTOC t

CTOC

lnðCTOC Þ ¼

0

Z 0

t

0

kTOC CO3 dt ¼  ln

CTOC t 0 ¼ kTOC CO3 t CTOC 0

ð14Þ

When the ozonation time was taken a constant for 25 min and lnðC TOC25 =C TOC 0 Þ values were plotted versus tCo3, Fig. 11 was ob0 tained. kTOC value was calculated as 2.11 M1 min1 from slope of plot. Table 3 also shows the experimental and calculated TOC removal values for different initial BPA concentrations after 25 min ozonation. As shown in Table 3, the calculated and experimental TOC values were almost similar for 0.509 mM BPA concentration. However, % relative error increased when the initial BPA concentration was decreased. 4. Conclusion

4

P/kHx10

a¼

0

1

2

3

4

5

4

[BPA]x10

Fig. 10. The change of P/kH ratios towards initial BPA concentrations.

In this study, the optimum decomposition pH was determined as 3.0 for BPA ozonation and the stoichiometric ratio between ozone and BPA was calculated to be 10.30 at this pH. Ten different intermediates formed during ozonation of BPA were identified by GC–MS and LC–MS/MS. The presence of malonic and oxalic acids after an ozonation time of 5 min demonstrated that the mineralization of intermediates began simultaneously with the ozonation of BPA. The first-order rate constants were calculated in the range of 19.3–13.3 s1. Similarly, the value of the second order rate constant determined in a range of 2.18  104–3.56  104 M1 s1. When BPA concentration was raised from 0 to 0.509 mM, Henry’s constant, kH, increased 5.5% depending on growth of ionic strength of BPA solution during ozonation. In this study, mineralization of BPA solution was not as successful as degradation of BPA for

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E. Kusvuran, D. Yildirim / Chemical Engineering Journal 220 (2013) 6–14

Table 3 The experimental and calculated TOC25/TOC0 values of BPA for different initial BPA concentrations. Initial BPA (mmol L1)

Consumed O3 by BPA (mmol L1)

TOC0 (mg L1)

TOC25 (mg L1)

%TOC25/TOC0 experimental

%TOC25/TOC0 calculated

Relative error (%)

0.509 0.381 0.254 0.127 0.051

0.161 0.117 0.110 0.103 0.095

91.4 68.6 45.0 22.9 9.2

68.8 47.7 30.3 15.1 6.0

75.2 69.6 67.3 66.2 65.0

71.2 78.1 79.1 80.5 81.9

5 12 18 22 26

0

ozonation times up to 25 min. kTOC value was estimated as 2.11 M1 min1. At the end of 25 min ozonation time, all BPA molecules were completely degraded, while BPA were mineralized only about 30%.

References [1] C.A. Staples, P.B. Dome, G.M. Klecka, S.T. Oblock, L.R. Harris, A review of the environmental fate, effects, and exposures of bisphenol A, Chemosphere 36 (1998) 2149–2173. [2] H. Okada, T. Tokunaga, X. Liu, S. Takayanagi, A. Matsushima, Y. Shimohigashi, Direct evidence revealing structural elements essential for the high binding ability of bisphenol A to human estrogen-related receptor-gamma, Environ. Health Perspect. 116 (2008) 32–38. [3] R.A. Keri, S.-M. Hob, P.A. Hunt, K.E. Knudsen, A.M. Soto, G.S. Prins, An evaluation of evidence for the carcinogenic activity of bisphenol A, Reprod. Toxicol. 24 (2007) 240–252. [4] H.J. Lee, S. Chattopadhyay, E.Y. Gong, R.S. Ahn, K. Lee, Antiandrogenic effects of bisphenol a and nonylphenol on the function of androgen receptor, Toxicol. Sci. 75 (2003) 40–46. [5] A.M. Soto, L.N. Vandenberg, M.V. Maffini, C. Sonnenschein, Does breast cancer start in the womb, Basic Clin. Pharmacol. Toxicol. 102 (2008) 125–133. [6] EFSA, Opinion of the scientific panel on food additives, flavourings, processing aids and materials in contact with food on a request from the commission related to 2,2-bis(4-hydroxyphenyl)propane (bisphenol A), EFSA J. 48 (2006) 1–75. [7] A.M. Calafat, Z. Kuklenyik, J.A. Reidy, S.P. Caudill, J. Ekong, L.L. Needham, Urinary concentrations of bisphenol A and 4-nonylphenol in a human reference population, Environ. Health Perspect. 113 (2005) 391–395. [8] V.K. Gupta, I. Ali, V.K. Saini, Removal of chlorophenols from wastewater using red mud: an aluminum industry waste, Environ. Sci. Technol. 38 (2004) 4012– 4018. [9] A.K. Jain, V.K. Gupta, S.J. Suhas, Removal of chlorophenols using industrial wastes, Environ. Sci. Technol. 38 (2004) 1195–1200. [10] M. Neamtu, F.H. Frimmel, Degradation of endocrine disrupting bisphenol A by 254 nm irradiation in different water matrices and effect on yeast cells, Water Res. 40 (2006) 3745–3750. [11] Z. Guo, R. Feng, Ultrasonic irradiation-induced degradation of lowconcentration bisphenol A in aqueous solution, J. Hazard. Mater. 163 (2009) 855–860. [12] M.F. Brugnera, K. Rajeshwar, J.C. Cardoso, M.V.B. Zanoni, Bisphenol A removal from wastewater using self-organized TIO2 nanotubular array electrodes, Chemosphere 78 (2010) 569–575. [13] P. Zhang, G. Zhang, J. Dong, M. Fan, G. Zeng, Bisphenol A oxidative removal by ferrate (Fe(VI)) under a weak acidic condition, Sep. Purif. Technol. 84 (2012) 46–51. [14] R. Huang, Z. Fang, X. Yan, W. Cheng, Heterogeneous sono-Fenton catalytic degradation of bisphenol A by Fe3O4 magnetic nanoparticles under neutral condition, Chem. Eng. J. 197 (2012) 242–249. [15] K. Ceulemans, S. Compernolle, J. Peeters, J.-F. Müller, Evaluation of a detailed model of secondary organic aerosol formation from a-pinene against dark ozonolysis experiments, Atmos. Environ. 44 (2010) 54345442. [16] Y. Lee, U.V. Gunten, Quantitative structure-activity relationships (QSARs) for the transformation of organic micropollutants during oxidative water treatment, Water Res. 46 (2012) 6177–6195. [17] I.A. Alaton, I.A. Balcioglu, D.W. Bahnemann, Advanced oxidation of a reactive dyebath effluent: comparison of O3, H2O2/UV-C and TiO2/UV-A processes, Water Res. 36 (2002) 1143–1154. [18] A. Colombo, G. Cappelletti, S. Ardizzone, I. Biraghi, C.L. Bianchi, D. Meroni, C. Pirola, F. Spadavecchia, Bisphenol A endocrine disruptor complete degradation using TiO2 photocatalysis with ozone, Environ. Chem. Lett. 10 (2012) 55–60. [19] E. Kusvuran, D. Yildirim, F. Mavruk, M. Ceyhan, Removal of chlorpyrifos ethyl, tetradifon and chlorothalonil pesticide residues from citrus by using ozone, J. Hazard. Mater. 241–242 (2012) 287–300. [20] N. Masomboon, C. Ratanatamskul, M.-C. Lu, Kinetics of 2,6-dimethylaniline oxidation by various Fenton processes, J. Hazard. Mater. 192 (2011) 347–353.

[21] E. Kusvuran, A. Samil, O.M. Atanur, O. Erbatur, Photocatalytic degradation kinetics of di- and tri-substituted phenolic compounds in aqueous solution by TiO2/UV, Appl. Catal. B 58 (2005) 211–216. [22] Sß . Gül, Ö. Özcan-Yıldırım, Degradation of Reactive Red 194 and Reactive Yellow 145 azo dyes by O3 and H2O2/UV-C processes, Chem. Eng. J. 155 (2009) 684– 690. [23] A. Ozlem Yıldırım, Sermin Gul, Orkide Eren, Erdal Kusvuran, A comparative study of ozonation, homogeneous catalytic ozonation, and photocatalytic ozonation for C.I. Reactive Red 194 azo dye degradation, Clean: Soil, Air, Water 39 (2011) 795–805. [24] H. Angermann, K. Wolke, C. Gottschalk, A. Moldovan, M. Roczen, J. Fittkau, M. Zimmer, J. Rentsch, Electronic interface properties of silicon substrates after ozone based wet-chemical oxidation studied by SPV measurements, Appl. Surf. Sci. 258 (2012) 8387–8396. [25] C.H. Wu, Decolorization of C.I. Reactive Red 2 in O3, Fenton-like and O3/Fentonlike hybrid systems, Dyes Pigment 76 (2008) 187–194. [26] J. Lee, H. Park, J. Yoon, Ozonation characteristics of bisphenol A in water, Environ. Technol. 24 (2003) 241–248. [27] M. Deborde, S. Rabouan, J.P. Duguet, B. Legube, Kinetics of aqueous ozoneinduced oxidation of some endocrine disruptors, Environ. Sci. Technol. 39 (2005) 6086–6092. [28] S. Irmak, O. Erbatur, A. Akgerman, Degradation of 17-estradiol and bisphenol A in aqueous medium by using ozone and ozone/UV techniques, J. Hazard. Mater. B126 (2005) 54–62. [29] F.J. Rivas, Á. Encinas, B. Acedo, F.J. Beltrán, Mineralization of bisphenol A by advanced oxidation processes, J. Chem. Technol. Biotechnol. 84 (2009) 589– 594. [30] T. Garoma, S. Matsumoto, Ozonation of aqueous solution containing bisphenol A: effect of operational parameters, J. Hazard. Mater. 167 (2009) 1185–1191. [31] S. Rodriguez-Mozaz, M.J. López de Alda, D. Barceló, Monitoring of estrogens, pesticides and bisphenol A in natural waters and drinking water treatment plants by solid-phase extraction–liquid chromatography–mass spectrometry, J. Chromatogr., A 1045 (2004) 85–92. [32] M. Deborde, S. Rabouan, P. Mazellier, J-P. Duguet, B. Legube, Oxidation of bisphenol A by ozone in aqueous solution, Water Res. 42 (2008) 4299–4308. [33] L. Alder, K Greulich, G. Kempe, B. Vieth, Residue analysis of 500 high priority pesticides: better by GC–MS or LC–MS/MS?, Mass Spectrom Rev. 25 (2006) 838–865. [34] IOA Standardisation Committee – Europe, Iodometric method for the determination of ozone in a process gas, International Ozone Association, Brussels, Revised Standardized Procedure 001/96, 1996. [35] E. Kusvuran, O. Gulnaz, A. Samil, M. Erbil, Detection of double bond-ozone stoichiometry by an iodimetric method during ozonation processes, J. Hazard. Mater. 175 (2010) 410–416. [36] R.Y. Peng, H.J. Fan, Ozonolytic kinetic order of dye decoloration in aqueous solution, Dyes Pigment 67 (2005) 153–159. [37] A. Lopez-Lopez, J.S. Pic, H. Debellefontaine, Ozonation of azo dye in a semibatch reactor: a determination of the molecular and radical contributions, Chemosphere 66 (2007) 2120–2126. [38] H. Tomiyasu, H. Fukutomi, G. Gordon, Kinetics and mechanism of ozone decomposition in basic aqueous solution, Inorg. Chem. 24 (1985) 2962–2966. [39] K. Sehested, H. Cotfltzen, J. Holcman, C.H. Flscher, E.J. Hart, The primary reaction in the decomposition of ozone in acidic aqueous solutions, Environ. Sci. Technol. 25 (1991) 1589–1596. [40] T.W.G. Solomons, C.B. Fryhle, Organic Chemistry, ninth ed., Wiley, USA, 2007. [41] R. Andreozzi, V. Caprio, I. Ermellino, A. Insola, V. Tufano, Ozone solubility in phosphate-buffered aqueous solutions: effect of temperature, tert-butyl alcohol, and pH, Ind. Eng. Chem. Res. 35 (1996) 1467–1471. [42] M.M. Huber, A. Göbel, A. Joss, N. Hermann, D. Löffler, C.S. Mcardell, A. Ried, H. Siegrist, T.A. Ternes, U. von Gunten, Oxidation of pharmaceuticals during ozonation of municipal wastewater effluents: a pilot study, Environ. Sci. Technol. 39 (2005) 4290–4299. [43] F.J. Beltrán, P. Pocostales, P.M. Álvarez, F. López-Pineiro, Catalysts to improve the abatement of sulfamethoxazole and the resulting organic carbon in water during ozonation, Appl. Catal., B 92 (2009) 262–270. [44] E. Kusvuran, O. Gulnaz, A. Samil, Ö. Yildirim, Decolorization of malachite green, decolorization kinetics and stoichiometry of ozone-malachite green and removal of antibacterial activity with ozonation processes, J. Hazard. Mater. 186 (2011) 133–143.