Electrochemical mineralization of chlorophenol by ruthenium oxide

Journal of the Taiwan Institute of Chemical Engineers 69 (2016) 106–117

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Electrochemical mineralization of chlorophenol by ruthenium oxide coated titanium electrode Rohit Chauhan, Vimal Chandra Srivastava∗, Ajay Devidas Hiwarkar Department of Chemical Engineering, Indian Institute of Technology Roorkee, Roorkee 247667, Uttarakhand, India

a r t i c l e

i n f o

Article history: Received 26 April 2016 Revised 1 October 2016 Accepted 7 October 2016 Available online 28 October 2016 Keywords: Chlorophenol Electrochemical degradation Specific energy consumption Current efficiency Kinetics Degradation mechanism

a b s t r a c t Electrochemical oxidation of 4-chlorophenol (CP) was investigated (in terms of chemical oxygen demand (COD) and CP removal efficiencies) by using a dimensionally stable anode (DSA) namely ruthenium oxide coated titanium (Ti/RuO2 ) electrode. Effect of process conditions such as current density (j), electrolyte concentration (m), initial pH (pHo ), time (t) and initial CP concentration (Co ) has been studied. Current efficiency (CE) and specific energy consumption (SEC) were also measured. Gas chromatograph-mass spectrometry (GC/MS) analysis was used to understand the CP mineralization mechanism which has been established on the basis of intermediates identified such as benzoquinone, hydroquinone and organic acids. Reaction kinetics was expressed by pseudo-first order kinetic model. Maximum COD removal efficiency of 96.7% and CP removal efficiency of 97.2%, respectively, was observed at j = 222.22 A/m2 , t = 180 min, pHo = 5.2 and m = 400 mg/l with SEC = 655 kWh/kg COD. Operating cost based on the studies performed on laboratory scale EC reactor has been calculated and compared with those reported for other pollutants degradation. © 2016 Taiwan Institute of Chemical Engineers. Published by Elsevier B.V. All rights reserved.

1. Introduction Phenols and its derivatives such as 4-chlorophenol (CP) are discharged by many industrial units such as oil refineries, petrochemical, plastic, pharmaceutical, and food-processing plants [1,2]. Among various phenolic compounds, CP is one of the most toxic phenolic compounds which can enter the human body by ingestion, inhalation or dermal adsorption. Considering its hazardous and toxic nature, it is necessary to degrade CP present in the water by suitable methods [3–5]. Physical, chemical and biological methods of wastewater treatment are generally not efficient in removing CP. Advance oxidation processes which include electrochemical (EC) processes such as electro-oxidation, electrofenton, electrocoagulation, etc. can be used for oxidizing many organic pollutants present in the wastewater [6–10]. Electro-oxidation process uses electron as oxidation reagent and converts organic pollutants into CO2 and H2 O [11]. Amount of sludge produced in the electro-oxidation process is negligible because of the oxidation of organics and it does not produce secondary pollution as well [12]. During electro-oxidation, pollutants present in the water get destroyed (when they get adsorbed on the electrode surface) by the electrons generated by the

∗

Corresponding author. Fax: +91 1332 276535. E-mail addresses: [email protected] (R. Chauhan), [email protected], [email protected] (V.C. Srivastava).

electrode. Also, oxidants generated in the solution mineralize the pollutants [13–17]. Behavior of EC oxidation depends upon the nature of electrode material i.e. which metal is used as a support for coating of metal oxides. Dimensionally stable anodes (DSA) are highly stable and do not corrode in the wastewater. These coated electrodes show better efficiency for oxidation of organic compounds. Several types of DSA like boron diamond doped (BDD) [6,11,18], Ti/RuO2 /IrO2 /TaO2 -coated titanium and graphite anodes [10], Ti/SnO2 and Ti/IrO2 anodes [14], titanium based DSA electrodes [19], platinum electrodes [20], PbO2 anode [21], etc. have been used in the literature for the treatment of various wastewaters. Ti/RuO2 is one such DSA electrode in which RuO2 is highly stable metal oxide even in strong acidic conditions and gives suitable oxidants which are used for oxidation of organic compounds [8,10,12,22]. Hou et al. [23] studied EC oxidation of bisphenol with the carbon aerogel electrode. Anand et al. [24] studied EC treatment of alkali decrement wastewater containing terephthalic acid using iron electrodes. Kumar et al. [12] investigated mechanism of EC oxidative degradation of p-nitro phenol by using Ti/RuO2 electrode. Azzam et al. [7] studied the EC oxidation of CP solutions using a DSA, which was made of pure titanium sheet mesh coated with Ti/TiO2 and RuO2 film. The influence of current density, pH and initial CP concentration within 2 h was investigated. In this study, only CP destruction efficiency was studied, however, its complete mineralization was not investigated. Moreover, specific energy consumptions and mechanism of destruction was not studied.

http://dx.doi.org/10.1016/j.jtice.2016.10.016 1876-1070/© 2016 Taiwan Institute of Chemical Engineers. Published by Elsevier B.V. All rights reserved.

R. Chauhan et al. / Journal of the Taiwan Institute of Chemical Engineers 69 (2016) 106–117

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Present study aimed to investigate mineralization of CP in terms of chemical oxygen demand (COD) and CP removal using Ti/RuO2 electrode under various operating conditions. Reaction mechanism of degradation of CP has been proposed on the basis of intermediates identified during the reaction with the help of sophisticated instruments.

propriate mass of CP (commercially purchased by Loba Chemie Pvt. Ltd., Mumbai, India) into distilled water. Ti/RuO2 electrodes were purchased from Titanium Tantalum Ltd. Company, Chennai, India.

2. Materials and methods

Electro-oxidation of CP was studied in a lab-scale glass batch reactor with circular cross section having 1 l volume and 11 cm diameter. It was equipped with two Ti/RuO2 electrodes as shown in Fig. 1. All the batch experiments were run with controlled action of current electrolysis process using a direct current (D.C.) source (4818A10). Configuration of electrodes was as follows: acquiring

2.1. Materials All chemicals used in the present study, were of analytical grade. Synthetic wastewater was prepared by dissolving ap-

2.2. Degradation of 4-chlorophenol using electrochemical method

Fig. 1. (a) Schematic diagram and (b) actual photograph of the experimental setup.

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equal thickness of 1.5 mm and dimension of 10 × 10 cm2 , however, practical surface area used was 90 cm2 for batch runs. Gap between anode and cathode was 1.0 cm. For easy stirring of aqueous solution at 30 0–40 0 rpm, base of electrodes was kept 6 cm from the bottom of the EC reactor. Voltage was varied in the range of 1–18 V and batch runs were conducted at room temperature of 25 ± 2 °C. Real photograph of the experimental setup is shown in Fig. 1.

2.3. Experimental procedure EC experiments were carried out using 1 l of CP solution filled into the reactor. Initial conductivity of the aqueous solution was adjusted by using NaCl as electrolyte. For adjustment of pH (measured using Elico Li-120, Chennai, India), dilute HCl and NaOH was used. Anode and cathode were connected to direct current power supply. EC treatment conditions were executed by varying the operating parameters in different range: current density (j) = 55.56–222.2 A/m2 , pH = 3.5–7.5, initial CP concentration (Co ) = 10 0–50 0 mg/l, electrolyte concentration (m) = 250–450 mg/l and time (t) = 180 min. Wastewater sample collected at the end of the experiment was centrifuged and filtered before determining the residual COD and CP concentration.

2.4. Theory Current efficiency (CE) is the ratio of actual charge required for oxidation of the compounds to the actual charge applied. It is estimated by following equation derived from Faraday’s law:

CE =

CODi − COD f 8It

F Vr

(1)

where CODi and CODf are the initial and final COD (g/l), F is the Faraday’s constant (96,485 C/mol), Vr is the reactor volume (l), I is the current intensity (A), ࢞t is the electrolysis time, and 8 is the oxygen equivalent mass (32 g O2 per mol/4e− ). Specific energy consumption (SEC) is the amount of energy absorbed (in kWh) per kg removal of COD:

SEC =

Vap It (CODi − COD f )Vr

(2)

where Vap is the applied voltage (V). Removal efficiency of COD and CP was calculated from the following relationship:

COD removal efficiency (% ) =

CODi − COD f × 100 CODi

(3)

Fig. 2. Effect of current density on (a) COD, (b) CP removal efficiency, and (c) SEC (kWh/kg COD) at pH 5.2, Co = 100 mg/l, electrolyte concentration m = 400 mg/l and at different time intervals.


Ci − C f CP removal/destruction efficiency (% ) = × 100 Ci

(4)

where CODi and CODf are the initial and final COD (mg/l) and Ci and Cf are the initial and final CP concentration (mg/l). 2.5. Instruments, analytic measurement and electrode characterization COD was measured with the help of COD analyzer (HACH/DR 50 0 0) and digestion unit (AL 38 SC, Aqualytic). The initial and residual CP concentration was determined by UVspectrophotometer (UV-1800, SHIMADZU, Serial No.-A114548) with maximum wavelength (λmax ) of 280 nm for CP. Concentration was calculated with help of calibration curve in between absorbance and concentration. Identification of intermediates was done using gas chromatograph mass spectra (GC-MS) analysis. For GC-MS analysis, 10 ml sample was mixed with 20 ml of di-chloro-methane (DCM) so as to extract CP and intermediates into DCM. This action has done 2– 3 times, after this a dehydrating agent sodium sulfate (anhydrous) was used to dehydrate the sample and concentrate it to about 5 ml and kept it at 4 °C. GC-MS used had a GC equipped with a capillary column (DB-5 ms, 30 × 0.25 mm, 0.25 m) as well as to a mass spectrometer (MS, Saturn 2100T) which was equipped with electron

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ionization (EI) source. Following program was used for identification: solvent delay of 180 s, split ratio of 15:1, range of scan from 50 to 500 at 3 scan/s, temperature of oven was increased from 50 °C (60 s) to 300 °C (60 s) whereas rate of ramp was 8 °C/min. For identification of species, NIST library was used. The species which match better than 60% were used for proposing mechanism of CP degradation. X-ray diffraction analysis (XRD) pattern of fresh electrode was recorded using Bruker AXS, Diffractometer D8, Germany. Diffraction pattern was matched with library of the International Centre for Diffraction Data (ICDD). Field emission scanning electron microscope (FE-SEM)/energy-dispersed X-ray (EDX) analysis of electrodes before and after EC treatment was done using QUANTA, Model 200 FEG, USA. The elements analysis using this method has error of 5–10%. 3. Results and discussion 3.1. Effect of applied current density and time Current density (j) is one the main parameters which affects the EC degradation of organic pollutants. Effect of j on CP degradation was studied in range of 55.56–222.22 A/m2 . Variation of COD and CP removal efficiencies with time is shown in Fig. 2a and b. COD (55.4–89.7%) as well as CP (67.1–90.4%) removal efficiencies

Fig. 3. Effect of pH on (a) COD, (b) CP removal efficiency, (c) SEC (kWh/kg COD) at current density j = 166.67 A/m2 , Co = 100 mg/l, electrolyte concentration m = 400 mg/l and at different time intervals; (d) distribution of CP species as a function of pH.

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increased when j was increased from 55.56 to 166.67 A/m2 . For j = 222.22 A/m2 , COD and CP removal efficiencies were maximum at 96.7% and 97.2%, respectively (Fig. 2a and b). This is because of the higher rate of production of free radicals at higher j and t, which increases the CP mineralization rate. Also, cathodic molecular oxygen reduction at higher j produces more amount of H2 O2 which in turn increases the rate of production of electrons which increases the rate of CP mineralization [25,26]. Variation of SEC with different value of j and time is shown in Fig. 2c. SEC values increased with an increase in j and t. This may also be due to the conversion of CP to stable intermediates that resist further oxidation after certain time of treatment, thus, increasing the value of SEC. Passivation of electrodes due to formation of impermeable film during electrolysis and oxygen gas evolution at anode can also increase SEC [27]. SEC value at j = 222.22 A/m2 where maximum COD and CP removal efficiencies were observed was found to be 655.06 kWh/kg COD. 3.2. Effect of pH pH of the CP solution is likely to greatly influence the COD and CP removal efficiencies. Fig. 3a and b shows the effect of pH

(within the range of 3.5–7.5) at j = 166.67 A/m2 , m = 400 mg/l, and Co = 100 mg/l. COD and CP removal efficiencies which increased with an increase in the pH owing to the increase in hydroxyl radical (OH• ) concentration. For pH 5.2, COD and CP removal efficiencies were 89.7% and 90.4%, respectively, however at pH 6.5, respective removal efficiencies were maximum at 96.1% and 97.1%, respectively. Respective removal efficiencies were nearly the same at pH 7.5. Effect of pH on SEC is shown in Fig. 3c. For treatment time of 180 min, SEC was found to be the minimum for pH 5.2, though SEC was only marginally lower than that at pH 6.5 or 7.5. The speciation of CP in deionized water is presented in Fig. 3d. The pKa value of CP is 9.38 [2,28,29]. CP species are present in the water in the form of C6 H5 OH and C6 H5 O− . For CP, the ratio of C6 H4 ClOH to C6 H4 ClO− is ≈0.9999 at pH 5.2, ≈0.9986 at pH 6.5, and ≈0.9869 at pH 7.5. Thus, CP is present in the neutral form at pH < 7.5 and it can be easily attacked by the oxidizing species. At lower pH values, H+ ions neutralize the oxidizing species and thus the removal efficiencies are lower at pH 5.2 as compared to that at pH 6.5. CE (shown in inset figure of Fig. 3a) increased with an increase in pH. Kirk et al. [30] reported CE increase from 3% to 13% in the

Fig. 4. Effect of initial CP concentration on (a) COD, (b) CP removal efficiency, and (c) SEC (kWh/kg COD) at j = 166.67 A/m2 , electrolyte concentration m = 400 mg/l and at different time intervals.


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Fig. 5. Effect of electrolyte (NaCl) concentration on (a) COD, (b) CP removal efficiency, and (c) SEC (kWh/kg COD) at j = 166.67 A/m2 , Co = 100 mg/l and at different time intervals.

pH range of 2 to 11 during oxidation of aniline with PbO2 electrode. Panizza et al. [31] reported maximum CE value of 15% during removal of methylene blue with boron-doped diamond electrode. Awad and Abuzaid [32] reported the pH effect on the phenol degradation rate with the help of graphite electrodes; and removal rate was slightly greater at pH 0.2 than for pH 0.5. Ruparelia and Soni [33] reported maximum CE value of 9% during RB-5 dye degradation by Ti/ RuO2 –SnO2 –Sb2 O5 anode. 3.3. Effect of initial CP concentration Industrial wastewater contains different concentration of CP, therefore, it is important to study the effect of initial CP concentration (Co ) on its degradation during the EC oxidation. Fig. 4a and b shows the effect of Co on the COD and CP removal efficiencies for various treatment time. As Co was increased from 10 0 mg/l to 50 0 mg/l, COD and CP removal efficiencies decreased. For Co = 100 mg/l, COD removal efficiency increased from 11.4% to 89.1% with an increase in treatment time from 15 to 180 min, whereas the CP removal efficiency increased from 25.5% to 90.3% during the same treatment time with other operating condition being j = 166.67 A/m2 , pH = 6.5 and m = 400 mg/l. For Co = 200, 30 0, 40 0 and 50 0 mg/l, COD removal efficiencies were found to be 79.1%, 73.6%, 62.4% and 53%, respectively; whereas CP removal efficiencies were: 83%, 74.0%, 65.6% and 56%, respectively, for

treatment time of 150 min. Thus, removal efficiencies decreased with an increase in Co due to insufficient availability of OH• at higher Co . However, net amount of COD removed is higher at Co which improves the CE at high Co (Fig 4a inset). Fig. 4c shows variation of SEC with Co and time. As the Co increases, SEC decreases because with an increase of Co , gross value of COD removed increases although the COD removal efficiency decreases. As the Co increases, amount of OH• available for CP degradation decreases due to which treatment efficiencies decrease. Also, COD and CP removal rates during EC oxidation depend upon the surface area of electrodes. At high Co , treatment efficiency decreases as the transfer of higher amount of CP across the electrode surface becomes difficult for direct oxidation [34]. 3.4. Effect of electrolyte (NaCl) concentration As the amount of supporting electrolyte (NaCl) increases, COD and CP removal efficiencies also increase (Fig. 5a and b). When NaCl concentration was 250 mg/l, COD and CP removal efficiencies were 66.3% and 70.6%, respectively, however, when NaCl concentration was 400 mg/l, COD removal efficiency increased to 96.7% and CP removal efficiency increased to 97.2%. These efficiencies increased only marginally when NaCl concentration increased to 450 mg/l. SEC, shown in Fig. 5c, first increased with time up to 30 min, thereafter, it remained constant up to 150 min of time and

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further increased. This is because, most of the COD removal occurred between 30–150 min of treatment. Thus, the mineralization of CP was greatly affected by NaCl concentration. High amount of NaCl helps in easy transfer of electrons between the electrode and the pollutant by providing supporting medium in the form of active spies such as chlorine and OH• . Cl− generated by NaCl gets converted into active chlorine species by interaction with active hydroxyl species. Overall, physisorbed active species (OH• ) cause complete mineralization whereas chemisorbed active species undergo selective oxidation [19,35]. 3.5. Reaction kinetics for degradation of CP In the indirect EC oxidation process, the removal rate of COD and CP is directly proportional to the organic compound (pollutant) concentration and oxidants produced in the bulk solution. Accordingly, the kinetics of COD and CP degradation can be expressed by the pseudo first-order kinetic model as [12,36]:

−

d [C] = k[C] dt

(5)

Value of pseudo-first order rate constant (k) for COD and CP degradation was calculated by constructing graph between ln(Co /Ct ) versus time (min) at different current densities (j):

ln

[C]o = kt [C]t

(6)

where Co is initial COD or CP concentration in mg/l, Ct is COD or CP concentration in mg/l at time t, t is time (min) and k is pseudo first-order reaction rate constant (min−1 ). Figs. 6 and 7 show pseudo first-order kinetics plots based on the COD and CP removal, respectively. Values of pseudo-first order rate constant (k) and correlation coefficient (R2 ) for COD and CP degradation are listed in Table 1. For j values of 55.56, 83.33, 111.11, 166.67 and 222.22 A/m2 , value of k for COD removal was found to be 6 × 10−3 , 6.8 × 10−3 , 9.5 × 10−3 , 14.6 × 10−3 and 25.6 × 10−3 min−1 , respectively, whereas respective values for CP degradation were found to be 4.5 × 10−3 , 6.3 × 10−3 , 8.8 × 10−3 , 13.8 × 10−3 and 18.8 × 10−3 min−1 , respectively. Overall, for both COD and CP degradation, values of pseudo-first order rate constant increased as the value of j, pH, Co and m increased. Since, higher value of k represents higher treatment rate, therefore, overall mineralization of CP is favored at higher values of j, pH, Co and m. It is also observed that value of k also approximately quadrupled when the value of j was increased four times (from 55.56 to 222.22 A/m2 ). Similarly, value of k increased 3.75 times when NaCl concentration (m) was increased 1.8 times (from 250 to 450 mg/l), while value of k increased 2.64 times only when Co was increased

Fig. 6. Pseudo first order kinetics of reaction for COD removal efficiency; effect of (a) current density, (b) pH, (c) initial CP concentration, and (d) electrolyte concentration.


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Fig. 7. Pseudo first order kinetics of reaction for CP removal efficiency; effect of (a) current density, (b) pH, (c) initial CP concentration, and (d) electrolyte concentration.

five times (from 100 to 500 mg/l). Thus, NaCl concentration (m) seems to be the most important parameter followed by j, pH and Co . Canizares et al. [6] reported the EC oxidation of CP using a BDD anode and reported k value of 0.0619 min−1 . Zhang et al. [36] used pseudo-first order kinetics model for removal of CP with Ti/α -PbO2 /β -PbO2 electrode and reported k value of 0.013 min−1 . Oturan et al. [37] reported k values in the range of 0.187– 0.498 min−1 at different j values for COD removal during CP degradation using carbon felt cathode and a Pt anode. Polcaro et al. [38] used pseudo first-order kinetic models and reported k values of 1.33 min−1 and 2.5 min−1 at pH 4 and 7, respectively, for removal of CP using Ti/PbO2 . Wang et al. [39] also reported pseudo first-order kinetics for CP removal of with stainless steel anode coupled with ultrasound where k values varied from 0.10 × 10−3 to 0.195 × 10−3 min−1 in the pH range of 2.48–9.08.

Chemical disintegration of superior oxides generates oxygen and other intermediate oxidation species which cause direct oxidation of organic pollutant [42,43]. Mineralization of CP also occurs due to the active species like ozone, hydrogen peroxide and active chlorine active chlorine, ozone, hydroxide or hydrogen peroxides which develop during indirect EC oxidation and oxidize organic pollutants into CO2 and H2 O. Oxide coated catalytic electrodes (MOx ) cause following types of reaction [44,45] at anode:

MOx + H2 O → MOx + OH. + H+ + e−

(9)

Higher oxide is converted by interaction of oxygen which is already present in anode due to oxide coated and hydroxyl radicals.

MOx (OH. ) → MOx+1 + H+ + e−

(10)

Electrolyte (NaCl) and degradation of CP produces Cl− which reacts with MOx (HO• ) by following reaction:

3.6. Mechanism for mineralization of CP Degradation of organics by electro-oxidation process happens due to the active species generated by DSA [40,41]. Oxidation of water generates superoxide metal (MO) and heterogeneous hydroxyl radicals (M(OH• )) on the electrode surface via equations given below.

M + H2 O → M(OH ) + H+ + e−

(7)

M(OH ) → MO + H+ + e−

(8)

MOx (HO. ) + Cl− → MOx (O. Cl ) + H+ + e−

(11)

Cl−

Anode oxide gives ion and oxygen on the surface of anode which may react with hypochloride radicals as given below:

MOx (O. Cl ) + Cl− → MOx+1 + H+ + e− MOx (O. Cl ) + Cl− → MOx +

1 1 O2 + Cl2 + e− 2 2

(12) (13)

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R. Chauhan et al. / Journal of the Taiwan Institute of Chemical Engineers 69 (2016) 106–117 Table 1 Kinetic rate constant for COD and CP removal with respect to different parameters. Parameter

COD

CP k1 (min−1 )

R2

j (A/m2 )

k (min−1 ) R2 Other condition: pH 5.2, Co = 100 mg/l, m = 400 mg/l

55.56 83.33 111.11 166.67 222.22

5.3 × 10−3 6 × 10−3 8.7 × 10−3 14 × 10−3 21.6 × 10−3

7.6 × 10−3 9.0 × 10−3 11.5 × 10−3 14.6 × 10−3 21.7 × 10−3

0.9914 0.9876 0.9828 0.9613 0.9318

0.9521 0.9489 0.9728 0.9689 0.9579

pH

Other condition: j = 166.67 A/m2 , Co = 100 mg/l, m = 400 mg/l

3.5 4.5 5.5 6.5 7.5

8.1 × 10−3 10.7 × 10−3 14 × 10−3 18.8 × 10−3 20.1 × 10−3

0.9509 0.9691 0.9689 0.9598 0.9669

Co (mg/l)

Other condition: j = 166.67 A/m2 , pH 5.2, m = 400 mg/l

100 200 300 400 500

5.3 × 10−3 6.2 × 10−3 8.6 × 10−3 10.2 × 10−3 14 × 10−3

m (mg/l)

Other condition: j = 166.67 A/m2 , pH 5.2, Co = 100 mg/l

250 300 350 400 450

7 × 10−3 9.3 × 10−3 14.1 × 10−3 21.6 × 10−3 24.3 × 10−3

0.9557 0.9661 0.9589 0.9638 0.9689

0.9686 0.9673 0.977 0.9579 0.96

10.5 × 10−3 12.2 × 10−3 13.8 × 10−3 17.6 × 10−3 18.2 × 10−3

0.9561 0.9664 0.9678 0.9571 0.9592

4.6 × 10−3 6.3 × 10−3 7.8 × 10−3 10.3 × 10−3 13.8 × 10−3

0.9729 0.9671 0.97 0.9639 0.9678

8.1 × 10−3 10.7 × 10−3 13.7 × 10−3 21.8 × 10−3 25.7 × 10−3

0.9836 0.9814 0.9774 0.9083 0.9116

Many intermediates were identified during GC-MS analysis of sample which was taken during EC degradation of CP. There are different ways of degradation of CP as shown in the proposed degradation mechanism shown in Fig. 8 which has been drawn on the basis of intermediates identified by GC-MS analysis and previous studies reported on CP degradation [3–5,7,35,39]. OH• radical may strike hydroxyl and chloro group on CP [3,39], which gives free radicals 4-chloro-p-benzoquinone. There are three possible routes of CP degradation. In the first possible path, CP undergoes dechlorination into phenol which after dehydrogenation gives hydroquinone which further gets converted into benzoquinone. Hydroxyl free radicals have enough energy so as to break the aromatic rings and thus oxidize the aromatic compounds. In the other two possible routes, first the ring breaks, however, the chlorine is still attached to the possible intermediates. In these possible CP degradation routes (II and III), 4-chloro-1,2-benzenediol and 4chloro-1,3-benzenediol, are the major intermediates. Further degradation causes formation of maleic acid which further degrades to form aliphatic carboxylic acids like acrylic acid, oxalic acid and succinic acid. In the next steps, OH• radicals convert the acrylic and succinic acid into malonic acid and acetic acid, however, oxalic acid gets converted into formic acid. In the last, all compounds get mineralized into small products i.e. CO2 and H2 O. Life of the oxidizing species and their oxidation potential greatly affects the decomposition of organic pollutants into intermediate aliphatic acids, CO2 and H2 O [35,46–51]. 3.7. Electrode characterization and operating cost analysis Fig. 9a shows the XRD analysis of fresh Ti/RuO2 electrode. XRD analysis showed the presence of anatase phased TiO2 (ICDD number 00-021-1272), and Ti2 O3 (ICDD number 0 0-010-0 063). Fig. 9b and c shows FE-SEM images of anode which shows surface structure of anode before EC treatment and after ≈50 number of EC runs. Ti/RuO2 electrode after its usage doesn’t show much change and still has uniform surface after use (Fig. 9c). Presence of RuO2 was confirmed by EDX analysis which showed presence

Fig. 8. Degradation mechanism of CP.


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Fig. 9. (a) XRD of fresh Ti/RuO2 electrode; FESEM images of electrode (b) before EC treatment, (c) used Ti/RuO2 electrode; EDX analysis of electrode (d) before EC treatment, (e) used Ti/RuO2 electrode after all EC degradation studied were performed.

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of Ru in the electrode both before and after EC treatment (Fig. 9d and e). It may be noted that elements Cl and Na were also found to be present on the electrode surface. Electrical energy cost (Penergy ) and electrode cost (Pelectrode ) are the two main factors which define expenditure of the system for treatment of 1 m3 of wastewater by EC method. COD and CP removal efficiencies were found to be 96.7% and 97.2%, respectively, at operation condition of j = 222.22 A/m2 , t = 180 min, Co = 100 mg/l, pHo = 5.2 and m = 400 mg/l. Consumption of electrical energy was calculated as ∼116.6 kWh/m3 when energy consumption was 655.06 kWh/kg COD. In India, cost of electrical energy is $0.05/kWh, therefore, Penergy , outcomes after calculation is $5.8 per m3 . Present electrode used was found to be highly stable; and that the electrode became unusable only after 60 runs treating 1 l of wastewater in each experiment (Fig. 9c). One electrode cost is approximately $11, therefore, Pelectrode becomes $183.3 per m3 . Total cost of the treatment process which is combination of Penergy and Pelectrode becomes $189.1 per m3 . This measurement of operating cost is basically only an approximation for the treatment of CP by EC degradation in aqueous solution. Kushwaha et al. [48] treated dairy industrial wastewater for organics removal by EC treatment method using iron electrode and total operating cost of the process was calculated to be the in range of $0.072–2.55 per m3 . Kumar et al. [12] calculated total operating expenditure for the EC degradation of PNP by Ti/RuO2 electrode as $226.7 per m3 . Sahu et al. [49] calculated operating cost of 248$/m3 for the treatment of sugar industry wastewater by EC degradation using aluminum electrode. 4. Conclusion This study explains EC treatment of 4-cholorophenol (CP) by Ti/RuO2 , a DSA electrode. Maximum removal efficiencies (COD ≈ 97% and CP ≈ 97%) were found to be for current density of 222.22 A/m2 , 180 min treatment time, electrolyte concentration of 400 mg/l with electrode spacing of 1.5 cm. Kinetics of COD and CP removal were found to follow pseudo-first-order kinetics. Mobile species like (Cl• and OH• ) generated on the surface of electrodes were found to play important role in mineralization of CP. Operating cost of the treatment process was found less than that of iron or aluminum electrodes reported in the literature. Moreover, EC oxidation of pollutants like CP by Ti/RuO2 electrode has the advantage of not producing any secondary solid waste such as scum. References [1] Suresh S, Srivastava VC, Mishra IM. Studies of adsorption kinetics and regeneration of aniline, phenol, 4-chlorophenol and 4-nitrophenol by activated carbon. Chem Ind Chem Eng Q 2013;19:195–212. [2] Thakur C, Dembla A, Srivastava VC, Mall ID. Removal of 4-chlorophenol in sequencing batch reactor with and without granular activated carbon. Desalin Water Treat 2014;52:4404–12. [3] Wang H, Wang J. Electrochemical degradation of 4-chlorophenol using a novel Pd/C gas-diffusion electrode. Appl Catal B Environ 2007;77:58–65. [4] Xu LJ, Wang JL. Degradation of chlorophenols using a novel Fe0/CeO2 composite. Appl Catal B Environ 2013;142–143:396–405. [5] Wang N, Li X, Wang Y, Quan X, Chen G. Evaluation of bias potential enhanced photocatalytic degradation of 4-chlorophenol with TiO2 nanotube fabricated by anodic oxidation method. Chem Eng J 2009;146:30–5. [6] Canizares P, Lobato J, Paz R, Rodrigo MA, Saez C. Electrochemical oxidation of phenolic wastes with boron-doped diamond anodes. Water Res 2005;39:2687–703. [7] Azzam MO, Al-Tarazi M, Tahboub Y. Anodic destruction of 4-chlorophenol solution. J Hazard Mater 20 0 0;75:99–113. [8] Yavuz Y, Koparal AS. Electrochemical oxidation of phenol in a parallel plate reactor using ruthenium mixed metal oxide electrode. J Hazard Mater 2006;136:296–302. [9] Kushwaha JP, Srivastava VC, Mall ID. Studies on electrochemical treatment of dairy wastewater using aluminium electrode. AIChE J 2011;57:2589–98. [10] Govindaraj M, Muthukumar M, Raju GB. Electrochemical oxidation of tannic acid contaminated wastewater by RuO2 /IrO2 /TaO2 -coated titanium and graphite anodes. Environ Technol 2010;31:1613–22.

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