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Research Article

Radiation Induced OH Free Radicals Degradation Process of Phenol in Aqueous Solutions – Environmental Implicationsa

Fathi Djouider1*

1Nuclear Engineering Department, Faculty of Engineering, King Abdulaziz University, Po Box 80204, Jeddah, 21589, Saudi Arabia

*Corresponding author Dr. Fathi Djouider, Nuclear Engineering Department, Faculty of Engineering, King Abdulaziz University, Po Box 80204, Jeddah, 21589, Saudi Arabia, Tel: +966 558822318;
Email: fdjouider@kau.edu.sa, fathid@yahoo.com

Submitted: 05-06-2015 Accepted: 07 -21-2015 Published: 08-07-2015

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Article

 

Abstract

The presence of the phenol, an organic refractory pollutant, in wastewater is one of the biggest environmental issues for both human health and life in aquatic ecosystems. In this study, gamma radiation induced degradation of phenol has been investigated. The radiolitically produced hydroxyl free radical OH• initiates the degradation via its addition to the phenol ring. After one hour, the phenol degradation was found 98% at a dose of 9 kGy. The initial major aromatic ring byproducts were identified as catechol, hydroquinone and benzoquinone. Further oxidation of these intermediates compounds via ring openings and bond breaking leads to a set of carboxylic acids and finally to a complete mineralization of phenol. The degradation rate decreased with increasing initial phenol concentration and followed first order kinetics. The influence of the pH on this degradation was also investigated. Increasing the pH from 6.6 to 12 decreased the phenol degradation efficiency from 95 to 73 %, respectively. Overall, the results showed that oxygen gas plays a key role in phenol degradation.

Keywords: Gamma Radiation; Hydroxyl Radical; Phenol Degradation; Oxidation; Radiation Induced

Introduction

A broad range of organic contaminants such as phenolic compounds (phenol, chlorophenols, and nitrophenol) are discharged in the water system from industrial, pharmaceutical and agricultural activities [1]. The US Environmental Protection Agency considers the contamination by these compounds of the hydric resources as a major environmental issue for both humans and life in aquatic ecosystems and prescribed a discharge limit in wastewater of less than 1 mg/L and a concentration of less than 1μg/L for drinking water [2]. Besides being highly irritating to the skin and eyes, phenol or its derivative contaminated drinking, cooking or bathing water can be one of the biggest risks to human health due to their toxicity to vital organs [3], mutagenic and carcinogenic effects [4]. Hence, removal of phenols from wastewater before its reuse or discharge to the environment is a vital issue.

Phenol is a colorless-to-white aromatic organic solid compound highly soluble in water. The hydroxyl group –OH is bound to the cyclic phenyl group –C6H5. Phenol and phenol compounds are known as refractory organic pollutants [5], which resist conventional biological degradation [6], ozonation [7], adsorption [8], solvent extraction [9] and electrochemical processes [10].

In the last few years, radical-based advanced oxidation processes proved significant importance for the alleviation of recalcitrant toxic pollutants in wastewaters. They involve the generation of some very reactive oxidizing species mainly the hydroxyl free radical (OH) which reacts unselectively at nearly diffusion control rates, ~ 109 mol dm-3 s-1, with aromatic hydrocarbons, unsaturated compounds and aliphatic alcohols and degrade them [11]. Oxidation of phenols and their derivatives by OH radicals generated by H2O2/Fe2+ Fenton and H2O2/Fe3+ Fenton-like reactions have been extensively studied [12]. This oxidation requires the presence of hydrogen peroxide and ferric or ferrous ions in the solutions to generate OH.

To the best of our knowledge, the radiation induced oxidation of phenols was seldom investigated. The objective of the present
laboratory study was to assess the advanced oxidation process of phenol with the gamma (γ) radiation induced OH. The effect of pH and initial phenol concentration on the removal percentage of phenol were investigated.

Materials and Methods

Phenol solutions of purity higher than 99 % was supplied by British Drug Houses (BDH) Chemicals and were used without any further purification. Phenol stock solutions of different concentrations were prepared using purified and deionized water by Millipore milli-Q system (resistivity ~ 18 MΩ-cm). In our work tert-butanol of analytical grade from BDH was used as received. The pH of solutions was measured with Electronic Instruments Ltd (EIL) 7200 glass electrode pH-meter calibrated before use. The deoxygenated solutions were obtained by bubbling phenol solutions with pure nitrous oxide gas (N2O) for around 15 min. The oxygenated solutions were prepared by bubbling with pure oxygen gas. The concentrations of phenol and its oxidation products were determined by the Hewlett- Packard 1090 high performance liquid chromatography (HPLC) equipped with a diode-array detector (DAD) and a reverse column Zorbax C18 (150 x 4.6 mm) column. The mobile phase in the reverse phase column consisted of a mixture of water and methanol (40:60, v/v) with an elution rate of 0.8 mL min-1. The eluate from the column was monitored at 280 nm. To monitor the removal of the phenol, irradiations were stopped periodically for the solution to be analyzed. All experiments were performed in a 100 cm3 Perspex beaker batch reactor at room temperature. The pH of the dilute phenol solution was found to be around 6.6. When necessary, the pH adjustments of the solutions were done by the addition of perchloric acid or borax.

Irradiation facility

Steady state irradiations were carried out using a 60Co γ-ray source (of activity 2000 Ci nominal activity) of average energy of 1.25MeV fixed at the end of a rod and shielded by a block of steel mounted on a concrete base for radiation protection purposes (Fig. 1). The source is moved to the irradiation position by pushing manually the rod in position “in”. A ruler was used to position the sample in the irradiation compartment in a determined distance from the source for a dose rate of 9 kGy/ hr which was used throughout this work. The design of the irradiation unit is such as the leaded access door remains locked when the rod is in the irradiation mode (Fig. 1a) and can be open only at the end of the irradiation time when the rod is pulled out (Fig. 1b).

envi fig 8.1
Figure 1: Top view of the irradiation compartment with the irradiated sample, (a): irradiation mode, (b): sample taken out.

Dosimetry

Dosimetry was performed using the dichromate chemical dosimeter at pH 9.2 [13]. It is based on measuring the amount of disappearance of chromate ion CrO4 2- at 370 nm, namely the net change in optical density due to irradiation. The dose is calculated according to the expression

where is the change in optical density of the chromate ion at 370 nm, is the reduction yield of this ion, is the density of the dosimeter solution, is the extinction coefficient of the chromate ion at 370 nm and is the cell optical pathlength. By taking = 0.20 μmol J-1, = 1 kg dm-3, = 4820 mol-1 dm3 cm-1 and = 1cm [13], the radiation dose in Gray is given by

Phenol solution irradiation

Irradiation of pure water with low energy transfer (LET) γ-radiation gives rise to transient species and stable byproducts homogeneously distributed in the bulk solution [11]

H2O _____> eaq, OH, H, H2, H2O2, H+, OH (3)

Table 1 gives the typical yields in μmol J-1 of these species and stable byproducts. A series of recombination reaction can occur between them as shown in Table 2.

envi table 8.1

Table 1: G-values of radiation induced species in pure water [11].

Table 2: Recombination reactions following pure water irradiation [11].

envi table 8.2

In dilute phenol solution the transient radicals react with phenol. The O2 − • and HO2 transient free radicals formed in the water radiolysis are much less reactive with phenol than OH , H and [14]. In the literature, the reported rate constant between O2 − • and phenol is 5.8 x 102 mol-1 dm3 s-1 [15]. , H and OH are competing for phenol as shown in Table 3.

Table 3: Reactions of phenol with OH , H and transient species [11].

envi table 8.3

Results and Discussions

As shown in Table 3, phenol is oxidized by hydroxyl radicals and reduced by hydrogen atom and the hydrated electrons. This results in its removal from the solution. The removal percentage at time of irradiation t, was calculated from the following equation

where is the initial phenol concentration (mol dm-3) and is the phenol concentration at time of irradiation t. The fractions of phenol reacting with OH, H and are respectively given by the expressions

eq 21

By using data from Table 1 and Table 3, these fractions are 94.3%, 5.4% and 0.3% respectively. This indicates that the hydroxyl radical is the primary species in the radiation induced degradation of phenol.

Phenol removal: effect of tert-butanol

Tert-butanol (2-methyl-2-propanol) is a good OH scavenger in aqueous solution [16]. The reaction proceeds by H abstraction
and yields the relatively unreactive radical • CH2C(CH3)2OH

OH + (CH3)3COH → CH2C(CH3)2OH + H2O (24)

k = 6 x 108 mol dm-3 s-1

Fig. 2 shows clearly that the degradation of phenol is greatly inhibited by the presence of this OH scavenger. For instance the degradation percentage was reduced from 95% to around 43% in the presence of 1 mM of tert-butanol at an absorbed dose of 4 kGy. In the presence of 1 mM of tert-butanol, around 40% of phenol was still detected after 1 hour of irradiation while in the absence of this scavenger all the phenol was degraded. This finding clearly indicates that hydroxyl radical is the primary degrading agent of phenol in aqueous solutions.

Figure 2: Effect of tert-butanol hydroxyl scavenger on phenol deg- radation: initial phenol concentration: 0.40 mM, dose rate: 9 kGy/hr. (█) without tert-butanol, (Δ ) with 0.01 M tert-butanol.

envi fig 8.2

Effect of gas bubbling

Fig. 3 shows that the radiation induced degradation of phenol in solution saturated with oxygen was increased nearly 9-fold when compared to N2O-saturated solution. When bubbled with N2O, hydrated electrons are quickly scavenged and converted into OH• [11]

e-aq+ N2O + H2O → OH + OH + N2 (25)

Although the yield of OH• increases in the N2O-saturated solution, G(OH)tot = G(OH) + G() = 0.28 + 0.28 = 0.56 μmol J-1, the degradation of phenol decreased but is not completely suppressed as small amount of radiolytically produced O2 builds up in the solution (reactions 10 − 13 in Table 2) and contribute to the oxidation of phenol with a low yield compared to the yield for O2-saturated solution. Sato and his co-workers [17] showed that the phenol decomposition yield was six-fold higher for γ-radiolysis in aerated solution than in deaerated one. The absence of oxygen inhibits the degradation of phenol and the buildup of its by-products.

envi fig 8.2

Figure 3: % Removal of phenol for different gases saturated solution. Initial phenol concentration: 0.40 mM, irradiation time: 1 hour, dose rate: 9 kGy/hr.


Kinetics of phenol degradation

Our experimental data for radiation induced removal of phenol in the absence of tert-butanol were found to fit a first order equation with respect to time as used previously by other authors [18]

         [c]=[co]e-kD           (26)

envi fig 8.3

         

where C is the remaining phenol concentration (mol L-1) at time t, is the initial concentration of phenol and k is the rate constant of the phenol removal (Gy-1). Taking the logarithm of its both sides, equation (26) can be linearized to:

eq 27

The rate constant of the phenol removal k was determined from linear plot of versus dose of irradiation and was found to be equal to 0.79 kGy-1 with a coefficient of correlation R2 = 0.9972 as shown in Fig. 4.

Figure 4: Plot of ln(C0/C) as function of absorbed dose. Initial phenol concentration: 0.40 mM, irradiation time: 1 hour, dose rate: 9 kGy/hr.

envi fig 8.4


The amount of degraded phenol per unit absorbed radiation energy, , calculated at each absorbed dose is expressed by [19]

eq 28

where is the change in concentration of the phenol (mol L-1) and is the absorbed dose (Gy). Table 4 shows that decreases with increasing absorbed dose. This is probably due to the competition between the phenol and its increasing byproducts (catechol and hydroquinone) for OH radicals as confirmed by previous work [20].

Table 4: Effect of dose on the degradation yield of phenol. Initial phenol concentration fixed at 0.40 mM.

envi table 8.4


Table 5 shows that for a fixed dose of 2.25 kGy, increases with increasing initial concentration of phenol although the removal percentage increased with the decrease in initial phenol concentration. The highest percentage phenol degradation was obtained when irradiating the solution with the lowest initial phenol concentration. Again this can be explained by the competition between the phenol and its byproducts for hydroxyl radicals.


Table 5: Variation of and the percentage removal of phenol with the initial phenol concentration. Dose of irradiation: 2.25 kGy.

envi table 8.5

The effect of the initial phenol concentration


To assess the effect of initial phenol concentration on its degradation, various initial phenol solutions of concentrations 0.1, 0.4, 0.8, 1.2 and 1.8 mM at near neutral pH, were irradiated at the same dose rate of 9 kGy/hr. As it is depicted in Fig. 5 and Table 6, the degradation of phenol decreases with increasing initial phenol concentration and was found to follow first order kinetics. The absorbed dose required to remove 90%, of the initial phenol concentration was computed using the following equation


Table 6: Dose and kinetic parameters obtained for the radiation induced degradation of phenol, pH = 6.6, dose rate: 9 kGy hr-1.

envi table 8.6

The experimental values of obtained from equation 29 were fit into the power function of the initial concentration with non linear least-squares method with the two parameters α and β as adjustable parameters. The values providing the best fit are 0.600 and -0.401 respectively with a coefficient of correlation R2 = 0.9844. The decrease of the removal efficiency of phenol with increasing its initial concentration is probably due to the increase in the concentration of the intermediates by-products which compete with the phenol for the radiation induced free radicals OH available in solution.

Figure 5: Initial phenol concentration vs. k profile for the radiation removal of 90% of initial phenol concentration: pH = 6.5, dose rate: 9 kGy hr-1.

envi fig 8.5

pH influence on the phenol degradation


The G-values of the transient species produced when irradiating water, the equilibrium OH / O• − and the degree of ionization of the phenol are all depending on the pH of the solution. Fig. 6 shows that the percentage degradation of phenol is higher at acidic than at alkaline pH. The higher rate of degradation was obtained at near neutral pH. These results are in agreements with some previous studies [21]. This pH dependence is probably related to the partially deprotonation of the hydroxyl group of the phenol in alkaline solution to form its conjugate base phenoxide ion [22]:


C6H5OH ⇌ C6H5O + H+ (pka = 10) (30)


and the hydroxyl radical abstracts hydrogen atom from the hydroxide ion to form the oxyl anion radical O• − [11]


OH + OH ⇌ O − + H2O pka = 11.9 (31)


Being nucleophilic species the O• − is less prone to attack the electron rich aromatic ring due to their lower redox potential vs SHE (1.73 V) compared to that of hydroxyl radical (2.73 V) [23].

 

Figure 6: Influence of pH initial concentration on the percentage degradation of phenol. Initial phenol concentration: 0.40 mM, dose rate: 9 kGy/hr.

envi fig 8.6

Mechanism of phenol degradation


Fig. 7 shows the formation and degradation of the three major by-products catechol, hydroquinone and p-benzoquinone. Being particularly unstable, [24], the o-benzoquinone was not detected in our experiments. Similar byproducts were also observed by some authors [25 - 27] using other advanced oxidation processes. This demonstrates that the very strong and non-selective electrophile free radical OH [28] attacks the ortho-position and the para-position of the phenol ring much faster than it will at other positions.


The –OH group attached to the phenol through the sp2 hybridized orbital of carbon atom, makes this aromatic compound highly reactive toward electrophilic addition because one of the oxygen lone pairs in the –OH group contributes to the electron density of the delocalized electron ring. Pulse radiolysis studies of the aqueous phenol solution have shown that the initial products formed when OH radical reacts with phenol are the OH-adducts ortho-dihydroxycyclohexadienyl and paradihydroxycyclohexadienyl transient intermediates [29]:


       C6H5(OH) + OH → C6H5(OH)2      (32)


This initial phenolic ring hydroxylation is in good agreement with the theoretical work done by Wu et al. [30] and Kiliç and his co-workers [31] who showed that the activation energies are in the order Eortho < Epara < Emeta which favors the orthoand para- addition over the meta- one. The para-dihydroxycyclohexadienyl and ortho-dihydroxycyclohexadienyl transient intermediates are then oxidized by either molecular oxygen O2 present in the solution (route 1 of Fig. 8) to yield hydroquinone and catechol, respectively with splitting off HO2 radicals.


C6H5(OH)2 + O2 → C6H4(OH)2 + HO2       (33)


or by OH• (route 2 of Fig. 8) to yield again hydroquinone and catechol

C6H5(OH)2 + OH → C6H4(OH)2 + H2O (34)

Figure 7: Formation and degradation of phenol byproducts during the radiation induced phenol oxidation of 0.4 mM, at dose rate 9 kGy/ hr. (█) catechol, (Δ) p-benzoquinone, (•) hydroquinone.

envi fig 8.7


Hydroquinone and catechol were also subsequently oxidized either by molecular oxygen O2 or by free radical OH• to p-benzoquinone and the unstable o-benzoquinone respectively [32] (Fig. 8). It has also been reported [33] that, depending on the pH of the solution, the hydroperoxyl free radical HO2 and its conjugate base superoxide radical O2 −•, although they are much less reactive with phenol than OH•, can further oxidize the hydroquinone and catechol to p-benzoquinone and o-benzoquinone respectively.

Figure 8: Reaction pathways for the radiation induced degradation of phenol.

envi fig 8.8

Figure 9: Proposed and simplified radiation induced degradation pathways of phenol: (1) phenol, (2) p- , (3) o-, (4) hydroquinone, (5) catechol, (6) p-benzoquinone, (7) o-benzoquinone, (8) maleic acid, (9) muconic acid, (10) glyoxylic acid, (11) malic acid, (12) oxalic acid, (13) formic acid, (14) succinic acid, (15) glycolic acid, (16) fumaric acid.

envi fig 8.9

These primary byproducts present maxima where their subsequent degradation starts by aromatic ring opening and C–C, C=C, C–O and C–H bond cleavages. This leads to numerous unsaturated carboxylic acids, as shown in Fig. 9, detected in traces and resulting in a slight decrease in the solution pH. They were identified as: maleic, fumaric, succinic, glyoxylic, oxalic, formic, acetic, muconic and malonic acids. These findings confirm the results of some other authors [34,35].


Conclusions


This work showed that the hydroxyl free radical OH initiates the degradation of aqueous phenol. This degradation depends on the absorbed dose and the pH of the solution and proceeds in three sequences

Phenol → OH-phenol adduct → ring products → aliphatic acids The initially high yield of hydroquinone and catechol supports the evidence that OH radical attacks the ortho- and para- position of phenol. The molecular oxygen present in solution plays a key role in this radiation induced degradation of phenol as it degrades the OH-phenol adduct to catechol and hydroquinone and further oxidizes these latter into o-benzoquinone and p-benzoquinone respectively. The highest radiation induced degradation was found at near neutral pH and alkaline environment is less prone to induce phenol degradation. Oxidation of these byproducts leads to a series of potentially less harmful aliphatic acids which do not need further mitigation.

Acknowledgements


The author would like to thank M. Wajeed Subhan for his invaluable help with the experimental setup.

References

References

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10.Prout BJ, Metreweli C. Pulmonary Aspiration after Fibre- endoscopy of the Upper Gastrointestinal Tract. 1972, 4(5835): 269-271.

11.Marom EM, McAdams HP, Erasmus JJ, Goodman PC. The many faces of pulmonary aspiration.1999, 172(1): 121- 128.

12.Cotran RS, Kumar V, Robbins SL. Robbins Pathologic Basis of Disease. 4th ed. Philadelphia, 1989.

13.Tobias D, Bruce AR. Heart-Lung Transplantation In Situs Inversus Totalis. 2009, 88(3): 1002-1003.

14.Nam OO, Jang JS, Noh MH, Park JI, Kim HJ et al. A Case of Aspiration Pneumonia after Upper Gastrointestinal Endoscopy. 2014, 14(30): 215-218.

15.Cooper GS, Kou TD, Rex DK. Complications Following Colonoscopy with Anesthesia Assistance: A Population- Based Analysis. 2013, 173(7): 551-556.

16.Ng A, Smith G. Gastroesophageal reflux and aspiration of gastric contents in anaesthetic practice. 2001, 93(2): 494- 513.

17.Friedrich K, Beck S, Stremmel W, Sieg A, BNG study group. Respiratory Complications in Outpatient Endoscopy with Endoscopist-Directed Sedation. 2014, 23(3): 255-259. 

18.Prout BJ, Metreweli C. Pulmonary Aspiration after Fibre- endoscopy of the Upper Gastrointestinal Tract. 1972, 4(5835): 269-271.

19.Prout BJ, Metreweli C. Pulmonary Aspiration after Fibre- endoscopy of the Upper Gastrointestinal Tract. 1972, 4(5835): 269-271.

20.Prout BJ, Metreweli C. Pulmonary Aspiration after Fibre- endoscopy of the Upper Gastrointestinal Tract. 1972, 4(5835): 269-271.

21.Marom EM, McAdams HP, Erasmus JJ, Goodman PC. The many faces of pulmonary aspiration.1999, 172(1): 121- 128.

22.Cotran RS, Kumar V, Robbins SL. Robbins Pathologic Basis of Disease. 4th ed. Philadelphia, 1989.

23.Tobias D, Bruce AR. Heart-Lung Transplantation In Situs Inversus Totalis. 2009, 88(3): 1002-1003.

24.Nam OO, Jang JS, Noh MH, Park JI, Kim HJ et al. A Case of Aspiration Pneumonia after Upper Gastrointestinal Endoscopy. 2014, 14(30): 215-218.

25.Cooper GS, Kou TD, Rex DK. Complications Following Colonoscopy with Anesthesia Assistance: A Population- Based Analysis. 2013, 173(7): 551-556.

26.Ng A, Smith G. Gastroesophageal reflux and aspiration of gastric contents in anaesthetic practice. 2001, 93(2): 494- 513.

27.Friedrich K, Beck S, Stremmel W, Sieg A, BNG study group. Respiratory Complications in Outpatient Endoscopy with Endoscopist-Directed Sedation. 2014, 23(3): 255-259. 

28.Prout BJ, Metreweli C. Pulmonary Aspiration after Fibre- endoscopy of the Upper Gastrointestinal Tract. 1972, 4(5835): 269-271.

29.Prout BJ, Metreweli C. Pulmonary Aspiration after Fibre- endoscopy of the Upper Gastrointestinal Tract. 1972, 4(5835): 269-271.

30.Prout BJ, Metreweli C. Pulmonary Aspiration after Fibre- endoscopy of the Upper Gastrointestinal Tract. 1972, 4(5835): 269-271.

31.Marom EM, McAdams HP, Erasmus JJ, Goodman PC. The many faces of pulmonary aspiration.1999, 172(1): 121- 128.

32.Cotran RS, Kumar V, Robbins SL. Robbins Pathologic Basis of Disease. 4th ed. Philadelphia, 1989.

33.Tobias D, Bruce AR. Heart-Lung Transplantation In Situs Inversus Totalis. 2009, 88(3): 1002-1003.

34.Nam OO, Jang JS, Noh MH, Park JI, Kim HJ et al. A Case of Aspiration Pneumonia after Upper Gastrointestinal Endoscopy. 2014, 14(30): 215-218.

35.Cooper GS, Kou TD, Rex DK. Complications Following Colonoscopy with Anesthesia Assistance: A Population- Based Analysis. 2013, 173(7): 551-556.

Cite this article: Djouider F. Radiation Induced OH Free Radicals Degradation Process of Phenol in Aqueous Solutions – Environmental Implications. J J Environ Sci. 2015, 1(2): 008.

 

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