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RuO2 anode: Mechanism, oxidation products and degradation pathway
Electrochimica Acta 296 (2019) 856e866
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Electrochimica Acta journal homepage: www.elsevier.com/locate/electacta
Amoxicillin electro-catalytic oxidation using Ti/RuO2 anode: Mechanism, oxidation products and degradation pathway Ravneet Kaur, Jai Prakash Kushwaha*, Neetu Singh Chemical Engineering Department, Thapar Institute of Engineering and Technology, Patiala, Punjab, India
a r t i c l e i n f o
a b s t r a c t
Article history: Received 25 September 2018 Accepted 17 November 2018 Available online 19 November 2018
Present work investigates the application of electro-catalytic oxidation (EO) technique using dimensionally stable anode, titanium coated with ruthenium dioxide (Ti/RuO2), for abatement of amoxicillin trihydrate (AMT), a much commonly prescribed antibiotic detected in water and wastewater. AMT removal efficiency (%ARE) and TOC (Total organic carbon) removal efficiency (%TRE) were measured by varying process parameters such as initial pH, current density (i), initial AMT concentration (C0) and supporting electrolyte (NaCl) concentration (S0). Mineralization current efficiency (%MCE) and specific energy consumption (SEC) values were evaluated and compared for different values of i and S0. Furthermore, decay kinetics of AMT was studied by varying i and C0. Moreover, AMT degradation and mineralization mechanism was explored in detail. Additionally, a possible pathway of AMT degradation/ mineralization was proposed by identifying the intermediates formed during EO reactions using UPLC-QTOF-MS. Besides, economic feasibility of EO treatment method was analysed by calculating the operating cost. The optimum current density and initial pH were found to be 5.88 mA cm2 and 7.0, at which, 60% ARE and 48% TRE were achieved in 60 and 240 min of electrolysis, respectively. Mineralization current efficiency was observed decreasing from 11.77% to 7.67% with increasing i value. © 2018 Elsevier Ltd. All rights reserved.
Keywords: Electro-catalytic oxidation Amoxicillin trihydrate Ti/RuO2 Mineralization current efficiency Specific energy consumption Degradation pathway
1. Introduction Water pollution is one of the major concerns of mankind in the present era. Urbanization, industrialization and population growth has led to consumption of enormous gallons of water, which in turn leads to generation of hefty volumes of wastewater. A number of treatment technologies have been developed and implemented at production sites to eliminate detrimental pollutants from wastewater, in order to make it suitable to be discharged in to aquatic environments. Since the introduction of electrochemical advanced oxidation processes (EAOPs) in the late 1970s, they have fetched much of the attention of scientific groups worldwide for treatment of bio-refractory and toxic organic pollutants from water [1e4]. EAOPs are based on in situ generation of hydroxyl radical ( OH), a powerful oxidizing agent which unselectively destroys the organic content in wastewaters. Among different EAOPs, electro-oxidation (EO) method of water treatment has become quite popular for its application in industrial wastewater treatment and target pollutant
* Corresponding author. E-mail addresses: [email protected] (J.P. Kushwaha), neetu.singh1479@gmail. com (N. Singh). https://doi.org/10.1016/j.electacta.2018.11.114 0013-4686/© 2018 Elsevier Ltd. All rights reserved.
destruction [5,6]. Major advantages of EO are: 1) no sludge generation, hence environmentally compatible; 2) minimum or no addition of chemicals required; 3) versatile; 4) amenable to automation; 5) operates at mild conditions and simple equipment, hence safe; 6) robust as reaction can be easily terminated by switching off the power [7]. EO technique destroys organic pollutants by two different pathways, i.e., “direct oxidation” at the anode surface and “indirect oxidation” in the bulk solution by oxidizing agents generated on the anode surface [5,8]. Performance and efficiency of EO technique have been investigated for degradation of persistent organic pollutants, such as pharmaceutical compounds, from wastewater [9e15]. Much focus has been given on removal of antibiotics by EO [16e20], since the studies in different parts of globe have reported presence of antibiotics in surface water, ground water and effluent from wastewater treatment plants (WWTPs) in the range of few ng L1 to 100 mg L1 [21e27]. Conventional physico-chemical and biological treatment methods used in WWTPs are unable to remove antibiotics from wastewater because of their biorefractory nature, thus resulting in their accumulation in surface and ground water [21,28]. It is a problematic issue as research has revealed continuous rise in antibiotic-resistant bacteria, which is dangerous for both humans
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and animals, because at present, there are quite limited alternatives to available antibiotics [29e31]. Amoxicillin trihydrate (AMT) is a semi-synthetic broad spectrum b-lactam antibiotic. It is one of the most prescribed antibiotics for children, adults and animals, for a number of bacterial infections, such as pneumonia, middle ear infections, strep throat, skin infections and urinary tract infections. It is commonly detected in sewage treatment plant effluents and surface waters [22,32], as only 20e30% of AMT is metabolized in human and livestock body systems and rest is excreted [33]. There were very few studies available in literature dealing with AMT removal using electrochemical oxidation technique [34e37]. In an interesting study, Sopaj et al., compared the capability of different anode materials such as BDD (Boron doped diamond), Ti/ RuO2-IrO2 (DSA), Pt, PbO2, carbon-fiber, carbon-graphite and carbon-felt for AMT removal. Complete mineralization was attained using BDD in 6 h after applying 41.66 mA cm2 of current density. Nonetheless, for same reaction conditions, less than 25% mineralization was reported for DSA. Similarly, Ganiyu et al., compared the performance of BDD, Pt, DSA and Ti4O7 as anode materials, and carbon-felt as cathode material for EO treatment of AMT. Carbonfelt cathode assisted in formation of H2O2 from compressed air circulated in to the EO cell. Ti4O7 exhibited better results than Pt and DSA, but was found to be less effective than BDD. Frontistis et al., performed AMT oxidation for current density values of 30 and 50 mA cm2 by means of BDD anode, and proposed possible pathways of degradation of AMT by BDD. Another study on BDD application for AMT removal was reported by Panizza et al. [60] wherein complete mineralization by electro-Fenton process was reported. However, the requirement of highly acidic pH conditions (pH z 3.0) for Fenton's reaction is a major drawback of this process. Likewise, BDD anodes have proven their competence by being dimensionally stable and giving supreme performance for destruction of other varieties of antibiotics as well [38e42]. However, these are comparatively much more expensive, which restricts their usage and questions their application on large-scale [5]. In another remarkable study, Santos et al., prepared a dimensionally stable anode (DSA), Ti/Pt/SnO2-Sb2O4 and reported successful removal of AMT from wastewater. Nevertheless, ~50% TOC abatement was achieved after 24 h of EO at 30 mA cm2of current density. More recently, titanium coated with ruthenium dioxide (Ti/ RuO2) has emerged as a promising anode material for degradation of organic pollutants because of its low cost, dimensional stability, high efficiency in strong acidic conditions, high electro-catalytic activity for generation of chloro-oxidant species (Cl2, OCl, HOCl) and other strong oxidizing agents ( OH, H2O2, O3) [43e46]. Such strong mechanical and chemical properties along with high oxygen evolution overpotential (z1.9 V) and low cost, make Ti/RuO2 a better alternative to rest of the anode materials. The degradation of organic pollutant (R) by EO treatment technique using Ti/RuO2 electrodes can follow the route of direct oxidation at the anode surface and/or indirect oxidation in the bulk solution. Direct anodic oxidation takes place via direct electron transfer from anode to organic molecule (R) as shown in following reaction: Ti/RuO2 þ R / Ti/RuO2.(R)ad / nCO2 þ oxidation products þ yHþ þ ze
857
chemisorbed OH, it gets oxidised to form RO, which is also called electrochemical “conversion” (Eq. (5)) [47]. Further oxidation of RO would lead to formation of CO2, H2O and other degradation products and ions. Ti/RuO2 þ H2O / Ti/RuO2.( OH)ad þ Hþ þ e (Physisorption)
(2)
Ti/RuO2.( OH)ad / Ti/RuO3 þ Hþ þ e (Chemisorption)
(3)
Ti/RuO2.( OH)ad þ R / Ti/RuO2 þ nCO2 þ oxidation products þ yHþ þ ze
(4)
Ti/RuO3 þ R / Ti/RuO2 þ RO
(5)
However, the inevitable but undesirable reaction of oxygen evolution occurs simultaneously as shown in following reactions: Ti/RuO2.( OH)ad / Ti/RuO2 þ 1/2O2 þ Hþ þ e
(6)
Ti/RuO2þ1 / Ti/RuO2 þ 1/2O2 þ Hþ þ e
(7)
Despite the advantages, studies on application of Ti/RuO2 for removal of antibiotics from wastewater are very few [45,49]. The novelty of present study lies in application of Ti/RuO2 electrodes for degradation and mineralization of AMT by EO treatment methodology and generation of plausible degradation pathway using Ti/ RuO2 anodes. Effect of important operating parameters, such as current density (i), initial pH of synthetic wastewater, supporting electrolyte concentration (S0) and initial AMT concentration (C0) on AMT removal efficiency (%ARE) and TOC removal efficiency (%TRE) of Ti/RuO2 anode was studied in detail. AMT decay kinetics was studied by varying i (mA cm2) and C0 (mg L1). Besides, mineralization current efficiency (%MCE) and specific energy consumption (SEC) were evaluated. Furthermore, the economic feasibility of the EO treatment technique using Ti/RuO2 anode was checked by calculating the operating cost. Major transformation products were identified using UPLC-Q-TOF-MS and a plausible degradation pathway of AMT was proposed. 2. Experimental 2.1. Materials Fresh synthetic wastewater was prepared at room temperature for each electro-oxidation experiment. Predefined quantity of AMT antibiotic (DSM Sinochem Pharmaceuticals, Punjab, India) and NaCl (Loba Chemie Pvt. Ltd., Mumbai, India) were dissolved in 1.5 L of double distilled water by placing the mixture over a magnetic stirrer for 1 h. NaCl was added as supporting electrolyte to enhance the conductivity of the synthetic wastewater. Initial pH of wastewater was adjusted using 0.1 M HCl or 0.1 M NaOH. All the chemicals used in the procedure were of analytical grade. Dimensionally stable Ti/RuO2 electrodes were purchased from Titanium Tantalum Products Ltd. Company, Chennai, India. 2.2. Electrochemical treatment
(1)
Electrolysis of water produces active oxygen in the form of hydroxyl radical at the anode surface which can either get physisorbed on to anode surface (Eq. (2)) or can get chemisorbed into the oxide matrix of the Ti/RuO2 anode, thus resulting in formation of Ti/RuO3 (Eq. (3)). When the pollutant species (R) come in contact of the physisorbed active oxygen, it results in to complete “combustion” as shown in Eq. (4). However, when R reacts with
The electrooxidation experiments were performed in a cuboid shaped reactor (13 cm 13 cm 13 cm) made from acrylic sheet. Set of four 1.5 mm thick Ti/RuO2 rectangular electrodes (10 cm 8.5 cm) were connected in parallel in bipolar mode such that two alternate electrodes worked as anode and the other two as cathode. The inter-electrode gap was fixed as 1 cm. 1.5 L of freshly prepared synthetic wastewater was introduced into the reactor and continuously mixed throughout the experiment with the help of a
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magnetic stirrer. Electrodes were immersed in the wastewater such that the bottom of the reactor was 2.5 cm apart from lower edges of the electrodes, which ensured proper rotation of the magnetic bead. Current controlled conditions for electrolysis were maintained using a regulated D.C. power supply (DIGITECH, Roorkee, India, Model: 4818A10; 0e20 V, 0e5 A). Electrooxidation reactions began when the power supply was switched on. The set of electrodes was properly washed before performing each experiment. The schematic diagram the electrochemical set-up is shown somewhere else [45]. In order to study the degradation of AMT using Ti/RuO2 electrodes, a set of experiments was performed over a range of parameters initial pH ¼ 2.0e9.0, current density (i) ¼ 1.47e5.88 mA cm2 (corresponding to applied current intensity in the range of 0.25e1.00 A), initial AMT concentration (C0) ¼ 10e50 mg L1 and NaCl concentration (S0) ¼ 1.0e2.0 g L1. Residual AMT concentration and TOC content were analysed by withdrawing samples from the reactor at pre-set time intervals. The mineralization reaction of AMT involves its conversion to þ CO2 and inorganic ions such as SO2 4 and NH4 as shown in following equation: þ 2 C16H19N3O5S$3H2O þ 28H2O / 16CO2 þ 3NHþ 4 þ SO4 þ 69H þ 70e (8)
Mineralization current efficiency (MCE, in %), which is defined as ratio of charge used for mineralization of a pollutant to total charge passed during the electrolysis process, was calculated using Eq. (9).
MCE ð%Þ ¼
TOC 0 TOC f zFV R 4:32 107 nIt
100
(9)
where, TOC0 and TOCf are initial and final TOC concentration (mg L1), respectively, F is the Faraday's constant (96,485 C mol1), z is the number of electrons involved in oxidation of AMT (z ¼ 70 (Eq. (8)), VR is volume of the reactor (L), n is number of carbon atoms in the AMT molecule (n ¼ 16), I is the applied current intensity (A) and Dt is the electrolysis time (h). In order to homogenize units, a conversion factor of 4.32 107 (3600 s h1 12,000 mg mol1) was used in the above equation. The economic feasibility of the electrooxidation process using Ti/RuO2 electrodes was checked by evaluating the specific energy consumption (SEC) using Eq. (10). SEC is defined as the amount of energy consumed in kWh for reduction of 1 g of organic load from wastewater.
SEC kWhðg TOC removedÞ1 ¼
UIt TOC 0 TOC f
VR
(10)
where, U is average voltage of electrochemical cell (V) and t is the electrolysis time (h). The AMT decay by EO process at any instant of time (t) followed pseudo-first order reaction kinetics with rate constant kf (sec1).
2.3. Instrumentation and analytical measurements Initial pH of synthetic wastewater was measured using Thermo Scientific Orion 5 star pH meter. The residual AMT concentration was determined using UVevis spectrophotometer (Perkin Elmer, Lambda 35). A calibration curve for AMT was prepared at lmax ¼ 228 nm. %AMT removal efficiency (ARE) was calculated using the following equation:
ARE ð%Þ ¼
ðC 0 C t Þ 100 C0
(11)
where, C0 and Ct are initial and final AMT concentration (mg L1) respectively. The amount of Total organic carbon (TOC) in the solutions was calculated using TOC-VCPH, PC-controlled TOC Analyser, Shimadzu (oxidation performed at 680 C through platinum catalysed combustion). Mineralization degree of AMT wastewater was evaluated by examining %TOC removal efficiency (TRE) over a period of time by following relationship:
TRE ð%Þ ¼
TOC 0 TOC f TOC 0
100
(12)
AMT transformation products were analysed by ultraperformance liquid chromatography coupled with mass spectrometry, using UPLC-XEVO-G2-XS/QTOF-MS (Waters) equipped with BEH C18 column (100 mm 2.1 mm; 1.7 mm). The mobile phase contained 0.1% of formic acid in water (A) and acetonitrile (B). The flow rate was set as 0.3 mL min1. The initial chromatography method condition held the mobile phase (90% A and 10% B) constant for 2.5 min. Thereafter a linear gradient to 50% B up to 5.5 min was followed by a linear increase to 90% B, kept constant for next 2.5 min. Finally, the mobile phase was returned to initial conditions of 10% B for 2 min. The MS analyser was operated in positive ion mode across the range 100e1000 m/z with source and desolvation temperature set as 120 C and 350 C, respectively. Nitrogen was employed as a carrier gas. The capillary voltage and cone voltage were set as 2000 V and 30 V, respectively. Flow rate for cone gas and desolvation gas were 60 L h1 and 800 L h1, respectively. Fourier transform infrared (FT-IR) spectral study was conducted using Spectrum 2 (Perkin-Elmer) FT-IR spectrophotometer. The transmission spectra obtained was plotted against the wavenumber in the range of 4000e400 cm1. 3. Results and discussion 3.1. Effect of initial pH on degradation and mineralization of AMT Efficiency of electrooxidation process for abatement of organic pollutants is highly pH dependent. The nature of various oxidation species generated on the anode surface and in the bulk solution strongly depends on the pH of wastewater under treatment. Effect of initial pH (2e9) of the solution on %ARE and %TRE with time was studied at current density, i ¼ 5.88 mA cm2, S0 ¼ 2 g L1 and C0 ¼ 50 mg L1, and is shown in Fig. 1a and b, respectively. As shown in Fig. 1a, %ARE increased with an increase in pH of the solution from 2 to 7. After 120 min of EO process, 33.22% of AMT was degraded at pH ¼ 2, whereas, at pH ¼ 7, similar removal was achieved in just 10 min of EO. Almost 60% of AMT was degraded in 45 min of EO at pH ¼ 7. Thereafter no significance increase in %ARE was observed. 40.77%, 48.89% and 52.59% AMT was degraded in 120 min of electrolysis for pH values of 3, 4 and 5.5 (natural pH of synthetic wastewater), respectively. With further rise in pH (8 and 9), there was no improvement in removal efficiency. Similar trend was observed in case of %TRE values (Fig. 1b). In case of highly acidic medium (pH ¼ 2), 11.53% of TOC was removed in 120 min of EO. However, for the similar electrolysis time and conditions, a significant enhancement in %TRE was observed when pH was raised up to neutral level 7, as 41.21% of TOC content in synthetic wastewater got degraded. For pH values of 3, 4 and 5.5, 18.07%, 27.29% and 37.53% of TRE was achieved in 120 min, respectively. ~40% TRE was attained for pH values of 8 and 9. Therefore, no significant change in TRE was observed when pH was raised above 7.0.
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following reactions: Ti/RuO2 þ HOCl / Ti/RuO2.( OCl)ad þ Hþ þ e
(17)
Ti/RuO2.( OCl)ad / Ti/RuO3 þ Cl
(18)
The dominance of these species is highly pH dependent. Cl2 is predominant in highly acidic conditions (pH < 1), whereas, HOCl dominates in the range of 3 < pH < 5 and ClO in pH > 9 [4,47,53]. Oxidation of AMT in acidic conditions mainly occurs through Cl2, which has low oxidation potential, and chemisorbed hydroxyl radical, which are predominant in acidic conditions. Furthermore, dissolved chlorine gets reduced in acidic conditions, resulting in oxygen evolution in the solution (Eq. (19)), which resulted in low removal efficiency of EO system for AMT and TOC content [50]. 2Cl2 þ 2H2O / O2 þ 4HCl
(19)
Chemisorbed hydroxyl radicals lead to partial conversion of AMT molecules (Eq. (5)), hence resulting in quite low value of %TRE as compared to %ARE. After the pH is raised, hypochlorous acid becomes the major chloro-oxidant responsible for degradation of AMT in synthetic wastewater, whose oxidation potential is higher than that of Cl2. Thus, a significant increase in values of %ARE and % TRE was observed with increase in pH values up to 7. Moreover, in this pH range, physisorbed active oxygen and physisorbed oxychloro species play their part in AMT degradation and mineraliza tion (Eqs. (4), (17) and (18)). The increase in amount of OH radicals with increase in pH is also a key factor behind improved values of % ARE and %TRE [54]. As discussed above, %ARE and %TRE increased with the increase in initial pH of the wastewater up to pH 7.0, with no significant difference for higher pH values of 8 and 9. Thus, the optimum pH value for the present study was found to be 7 and further experiments were conducted at neutral pH only. 3.2. Effect of current density (i) on degradation and mineralization of AMT
Fig. 1. Effect of initial pH on (a) AMT removal efficiency, %ARE, and (b) TOC removal efficiency, %TRE (Other parameters: i ¼ 5.88 mA cm2, S0 ¼ 2 g L1, C0 ¼ 50 mg L1).
Ti/RuO2 electrodes exhibit high electrocatalytic activity for generation of reactive chlorine species which play an important role in indirect oxidation of organic pollutants [5,8,50]. Chlorine is electro-generated by direct oxidation of Cl ion on the anode surface as demonstrated in reactions (13) and (14) [50e52]. Reaction of chlorine with water results in formation of hypochlorous acid (HClO) in the bulk solution (Eq. (15)). Furthermore, deprotonation of HOCl generates hypochlorite ion (ClO) (Eq. (16)) Ti/RuO2 þ Cl / Ti/RuO2.( Cl)ad þ e
(13)
2Ti/RuO2.( Cl)ad / 2Ti/RuO2 þ Cl2
(14)
Cl2 þ H2O / HOCl þ Hþ þ Cl
(15)
HOCl þ OH / ClO⁻ þ H2O
(16)
The oxygen transfer to molecules of pollutant species can also take place through adsorbed oxy-chloro radicals as shown in
The effect of applied current density, i (1.47e5.88 mA cm2), on EO process parameters by Ti/RuO2 electrodes at pH ¼ 7, C0 ¼ 50 mg L1, S0 ¼ 2 g L1, was studied systematically. Fig. 2a shows the effect of different values of applied i (mA cm2) on %ARE values with electrolysis Significant increase in %ARE values was observed when i values were increased from 1.47 to 5.88 mA cm2. AMT removal improved from 9.77% at 30 min to 22.75% at 120 min for i ¼ 1.47 mA cm2. When EO was carried on further, 26.32% of AMT content got degraded in 150 min, with no further increase up to 180 min. Similarly, for i ¼ 2.94 mA cm2, equilibrium was attained in 150 min of EO with 38.44% ARE. However, for i values of 4.41 mA cm2 and 5.88 mA cm2, equilibrium stage was achieved in 75 min and 45 min, with 53.65% and 61% AMT removal, respectively. AMT removal rate got enhanced with the increase in applied current because the amount of oxidant species (chemisorbed and physisorbed OH, Cl2, HOCl, ClO) generated at the anode surface according to Eqs. (2) and (3), (13)e(18), increased substantially when applied current intensity was increased. However, not more than ~61% of ARE was attained for 5.88 mA cm2 of applied current density for present experimental conditions. This might be because of the inevitable reaction of oxygen evolution taking place at anode surface at higher values of current intensity (Eqs. (6) and (7)). Efficiency of EO for TOC abatement was tested in the same range of current density (1.47e5.88 mA cm2) for prolonged electrolysis time up to 240 min (Fig. 2b). 19.15%, 36.46%, 46.18% and 48.62% TRE was achieved for current densities of 1.47, 2.94, 4.41 and
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Fig. 2. Effect of applied current density, i (mA cm2) on (a) AMT removal efficiency, %ARE, (b) TOC removal efficiency, %TRE, and (c) Mineralization current efficiency, %MCE and Specific energy consumption, SEC (kWh (g TOC removed)1) (Other parameters: pH ¼ 7.0, S0 ¼ 2 g L1, C0 ¼ 50 mg L1).
5.88 mA cm2, respectively in 240 min of EO. Threefold escalation in %TRE was observed, when i was increased from 1.47 to 5.88 mA cm2. In neutral pH range, physisorbed active oxygen and HOCl play the key role in mineralization of organic species. The amount of these oxidants improved significantly with increase in applied current, hence resulting in substantial improvement in % TRE values. Comparative analysis of %MCE and SEC for different values of i after 120 min of EO is shown in Fig. 2c. For different values of current density, SEC increased from 0.228 to 0.558 kWh (g TOC removed)1 with an increase in i from 1.47 to 5.88 mA cm2. However, %MCE decreased from 11.77% to 7.67% for the same values of current density. Even though, TOC removal rate was enhanced with the increase in applied current density, still the energy consumption of EO system increased because of elevation in average EO cell potential, U from 3.1 V to 5.2 V, with increase in i from 1.47 to 5.88 mA cm2. As discussed above, the unescapable reaction of oxygen evolution (Eqs. (6) and (7)) is enhanced at higher current intensities which competitively consumes the extra electrons. Although, at higher i, the rate of generation of OH radicals is elevated, the unreacted radicals tend to form hydrogen peroxide and hydroperoxyl radical according to Eqs. (20) and (21) [55], whose oxidation power is much less as compared to that of hydroxyl radical [56]. 2 OH / H2O2
(20)
H2O2 þ OH / HO2 þ H2O
(21)
Aggravation of such parasitic reactions at higher current intensity causes the loss in mineralization efficiency of EO process, thus increasing the overall cost of treatment. 3.3. Effect of supporting electrolyte concentration For the present study, NaCl was employed as a supporting electrolyte for all EO experiments. It served two purposes: (1) enhancement of electrical conductivity of the synthetic wastewater to perform the electrolysis; (2) deliver active chlorine species as oxidizing agents for abatement of organic pollutant (AMT) present in synthetic wastewater. In order to study its effect on performance of EO process, a range of 0.5e2 g L1 of NaCl concentration (S0) in synthetic wastewater was selected. Other operating parameters were set as: i ¼ 5.88 mA cm2, pH ¼ 7, C0 ¼ 50 mg L1. As depicted in Fig. 3a and b, the increase in S0 resulted in increase in both %ARE and %TRE values. ARE value got enhanced from 52.51% to 60.88% after 120 min of electrolysis, when S0 was increased from 0.5 to 2 g L1 (Fig. 3a). For the same set of S0 values and electrolysis time, % TRE values also showed threefold improvement from 16.31% to 48.67% (Fig. 3b). The mediated oxidation of AMT was carried out by the oxy-chloro species generated at the Ti/RuO2 surface according to Eqs. (13)-(18). It was evident that the amount of these reactive chloro oxidants increased with the increase in concentration of NaCl in synthetic wastewater, thus enhancing the efficiency of EO for removal of AMT and TOC. Correlation of S0, %MCE and SEC is shown in Fig. 3c. %MCE increased with exponential decay in SEC (kWh (g TOC removed)1), when S0 (g L1) was increased. With increase in amount of NaCl, the
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Fig. 3. Effect of NaCl concentration, S0 (g L1) on (a) AMT removal efficiency, %ARE, (b) TOC removal efficiency, %TRE, and (c) Mineralization current efficiency, %MCE and Specific energy consumption, SEC (kWh (g TOC removed)1) (Other parameters: i ¼ 5.88 mA cm2, pH ¼ 7.0, C0 ¼ 50 mg L1).
conductivity of synthetic wastewater got significantly increased from 1.32 mS cm1 (S0 ¼ 0.5 g L1) to 4.08 mS cm1 (S0 ¼ 2 g L1), thus lowering the average cell potential (U) from 10.2 V (S0 ¼ 0.5 g L1) to 5.2 V (S0 ¼ 2 g L1), which in turn decreased the amount of energy consumed by the electrolytic system. Higher quantity of NaCl amplified the generation rate of chloro-oxidant species in wastewater, consequently increasing the TOC decay rate, and thus the mineralization efficiency of EO by Ti/RuO2 anodes. Therefore, the overall cost of EO treatment method decreased and the %MCE increased, when the amount of NaCl in synthetic wastewater was increased from 0.5 g L1 to 2 g L1. 3.4. Influence of initial AMT concentration and decay kinetics Degradation of AMT antibiotic by EO treatment technique at any instant of time (t) can be demonstrated by pseudo-first-order kinetic model given as below:
ln
C0 ¼ kt Ct
(22)
For any value of C0 and i, concentration of AMT in the wastewater first decreases with time and then attains an equilibrium value of Ce, leaving behind a fraction of initial AMT concentration, which could not be removed at experimental conditions. Therefore,
a more appropriate way for representing the above Eq. (22) is the modified form given below [57]:
C Rt ¼ C Re 1 ekf t
(23)
where, CRt and CRe (mg L1) are the amount of AMT removed at any time, t and at equilibrium conditions, respectively, and kf is pseudofirst order rate constant. Non-linear regression using Marquardt's percent standard deviation (MPSD) error function was performed by fitting the experimental data in following equation [58]:
vffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi !2ffi u n u 1 X C C Rt;i;exp Rt;i;cal MPSD ¼ 100t n p i¼1 CRt;i;exp
(24)
where, ‘n’ and ‘p’ are number of measurements and number of parameters in the model, respectively, and the subscripts ‘exp’ and ‘cal’ represent the experimental and calculated values of CRt. Kinetic study for degradation of AMT was performed by varying the initial concentration of AMT (C0) in the range of 10e50 mg L1 at i ¼ 5.88 mA cm2, pH ¼ 7.0, S0 ¼ 2 g L1 and also by varying current density in the range of 1.47e5.88 mA cm2 at same experimental conditions with C0 ¼ 50 mg L1. The kinetic data fitted to pseudo-first-order kinetic model by varying C0 and i are shown by
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solid line in Fig. 4a and b, respectively. The best fit values of rate constant, correlation coefficient and MPSD are shown in Table 1. The equilibrium condition was achieved in 30 min for C0 ¼ 10 mg L1, 60 min for C0 ¼ 20, 30 and 40 mg L1, and 90 min for C0 ¼ 50 mg L1. Hence, the time range for the kinetic study of AMT degradation was set accordingly as presented in Fig. 4a. As evident from data in Table 1, R2 values (~0.99) and low MPSD values justify the aptness of pseudo-first-order kinetic model for AMT degradation by EO using Ti/RuO2 electrodes. Furthermore, the proximity of CRt,exp and CRt,cal values reinforces the fitness of pseudo-first-order kinetic model for current study. The values of kf decreased from 2.36 103 to 1.09 103 sec1, with the increase in C0 from 10 to 50 mg L1 (Table 1). These results suggest that the degradation rate of AMT by EO at its lower concentration levels was faster than its higher concentration levels. The nature and amount of oxidation species ( OH, HOCl, ClO, H2O2, HO2, etc.) generated at the anode surface and in the bulk solution remains the same when a fixed current is supplied to the EO system. When substrate (AMT) concentration in wastewater is low, the
Table 1 Pseudo-first order kinetic parameters. Parameter
Value
kf 103 (sec1)
R2
MPSD
i (mA cm2)
1.47 2.94 4.41 5.88 10 20 30 40 50
0.16 0.17 0.63 1.09 2.36 1.98 1.29 1.17 1.09
0.997 0.997 0.998 0.996 0.997 0.994 0.999 0.992 0.996
9.28 12.17 6.46 6.88 3.56 10.53 4.34 5.04 6.88
C0 (mg L1)
oxidants attacking its molecules for degradation are in plenty, thus resulting in faster removal. Conversely, the ratio of substrate concentration and oxidants decreases for higher values of AMT concentration, causing the slower degradation rate, and thus lower kf values. When i was increased from 1.47 to 5.88 mA cm2, the value of kf also increased from 0.16 103 to 1.09 103 sec1 (Table 1). As discussed in section 3.2, the increase in value of applied current intensity amplifies the generation of oxidants ( OH, HOCl, ClO, H2O2, HO2, etc.) responsible for degradation of AMT molecules in synthetic wastewater. This in turn leads to faster degradation of AMT, and hence, higher values of pseudo-first-order rate constant.
3.5. Plausible AMT degradation pathway
Fig. 4. Kinetics of AMT degradation (a) Effect of initial AMT concentration, C0 (mg L1) (b) Effect of current density, i (mA cm2).
In order to derive the degradation scheme of AMT during EO process by Ti/RuO2 electrodes, synthetic wastewater with 200 mg L1 of AMT was taken in reactor for electrolysis at neutral pH, 5.88 mA cm2 of current density and 2 g L1 of NaCl. Samples after electrolysis reaction were collected for analysis by UPLC-QTOF-MS. The mass spectrum of initial sample before reaction (Fig. SM e 1a) shows peaks corresponding to AMT molecule and its fragmentation ions. The mass/charge ratio (m/z) of 366 corresponds to the protonated molecule of AMT [M þ H]þ. Additional peak at m/ z value of 349 represents the fragmentation ion AF 1 [M e NH3 þ H]þ of AMT molecule. The opening of b-lactam ring of AMT leads to formation of other two fragments AF 2 and AF 3, corresponding to m/z values of 208 and 160, respectively (Table 2). Besides, m/z value of 388 is attributable to the sodium adduct ion [M þ Na]þ. The retention time of AMT was 0.87 min (Fig. SM e 1b). The UPLC chromatogram of sample withdrawn after 15 min of EO reaction is shown in Fig. SM e 2. Three major AMT degradation products (ADPs) were identified, whose retention time, elemental composition and proposed structure is shown in Table 2. The plausible degradation/mineralization pathway of AMT by EO using Ti/RuO2 electrodes is depicted in Fig. 6. The reactions occurring during EO process are quite complex and involve a number of oxidant species, such as OH, HOCl and Cl2, which degrade the organic pollutants via direct or mediated oxidation. Based on the information gathered from UPLC-Q-TOFMS analysis of samples, a possible reaction pathway for AMT degradation was proposed. As shown in Fig. 6, loss of NH3 group, and attack of active chlorine leading to decarboxylation of AMT molecule, resulted in formation of a reaction intermediate ADP 1 (m/z ¼ 339). Subsequent hydroxylation of ADP 1 by active oxygen in the form of hydroxyl radicals, generated another degradation product, ADP 3 with m/z ¼ 325. The replacement of entire aromatic moiety of AMT molecule by active chlorine formed ADP 2 (m/ z ¼ 269). Alternative route for formation of ADP 2 could also be via direct attack of Cl ions on the AMT molecule, thus leading to decarboxylation, loss of aromatic moiety and amine group. Further
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863
Table 2 Major AMT degradation products (ADPs) identified by UPLC-Q-TOF-MS. Compound
Rt (min)
m/z [MþH]þ
Elemental composition
ADP 1
0.703
339
C15H15ClN2O3S
ADP 2
0.754
269
C8H10Cl2N2O2S
ADP 3
1.18
325
C13H12N2O6S
AMT
0.87
366
C16H19N3O5S
AF1
e
349
C16H16N2O5S
AF2
e
208
C10H11N2O3
AF3
e
160
C6H9NO2S
attack of oxidation species on degradation products would lead to þ formation of aliphatic acids and inorganic ions (SO 4 and NH4 ). Finally, mineralization occurred after oxidation of aliphatic acids to form CO2 and H2O. Mineralization efficiency of nearly 48% was achieved at optimum conditions of EO in 4 h of reaction, signifying the amount of carbon content in synthetic wastewater which got converted to CO2 (inorganic carbon). Degradation pathway of amoxicillin has also been proposed by Refs. [34,35] by employing BDD and Ti4O7 as anodes for electrooxidation treatment, respectively. Hydroxyl radical physisorbed on to the surface of these nonactive anodes played the key role in degradation and mineralization of the antibiotic in both the cases. However, in present study the role of chloro-oxidants was inevitable. Additionally, FT-IR spectral study of initial and final sample was also performed to analyse the effect of EO reactions (Fig. 5). The stretch extending from 3700 to 2750 cm1 is due to eOeH group and vibrations due to intermolecular hydrogen bonding. The peak at 3309.21 cm1 is due to amide bond (eNeH group) in AMT molecule [43,49]. An additional peak at 1636.24 cm1 is the characteristic absorption peak of eC]O bond of eCOOH group. The
Proposed Structure
further peak at 645.42 cm1 is part of the fingerprint region and is mainly due to carbonyl group and bending vibrations within the molecule. Identical peaks could be noticed in the FT-IR spectrum of final sample withdrawn after EO of synthetic wastewater comprising of 50 mg L1 of AMT and 2 g L1 of NaCl at applied current of 1 A and pH 7 after 120 min. However, the intensities of these peaks were significantly decreased as compared to those in spectrum of initial sample. The spectral peaks thus support the findings discussed in previous section which proved the conversion and combustion of AMT molecules via direct and mediated electrolysis by Ti/RuO2 electrodes using EO methodology. 3.6. Operating cost analysis Economic feasibility of any treatment technique is essential for its applicability at high scale. For a WWTP based on EO technique, operational costs would comprise of cost of electrical energy consumed, cost of electrodes, pumping and mixing cost, wages, etc. Amongst these, the cost of anode material and electrical energy are of much importance. Hence, to calculate the operating cost of EO
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Fig. 5. FT-IR spectra of synthetic wastewater before and after EO treatment (i ¼ 5.88 mA cm2, pH ¼ 7.0, S0 ¼ 2 g L1, C0 ¼ 50 mg L1).
Fig. 6. Proposed degradation pathway of AMT by EO treatment using Ti/RuO2 anodes.
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treatment of wastewater under present study, the total cost of electrical power consumed and that of anode material was calculated. At optimum conditions for EO treatment of AMT (i ¼ 5.88 mA cm2, pH ¼ 7.0, S0 ¼ 2 g L1), 0.404 kWh of electrical energy was consumed for removal of 1 g of TOC from wastewater in 1 h of EO. The average cost of electrical energy in India is INR 5 per kWh. Therefore, the price of electrical energy (CEE) comes out to be INR 2.02 (g of TOC removed)1. Ti/RuO2 electrodes (10 cm 8.5 cm; thickness: 1.5 mm) used in present study were procured for INR 2500 each. Two electrodes were employed as anodes for EO treatment of AMT. According to the manufacturer, the life of these electrodes was 2.5 years at optimum experimental conditions (i ¼ 5.88 mA cm2, pH ¼ 7.0, S0 ¼ 2 g L1). Using EO technique under optimum conditions, ~34% TOC was removed from wastewater comprising AMT in 1 h. The electrode cost (CEL) hence comes out to be INR 28.88 (g of TOC removed)1. Therefore, the total cost of EO treatment (CEE þ CEL) for removal of 1 g of TOC is INR 28.88 (~$ 0.44). 4. Conclusions Maximum removal efficiencies is terms of AMT and TOC were achieved at neutral pH of wastewater. Mineralization current efficiency decreased from 11.77% to 7.67% with increase in current density from 1.47 to 5.88 mA cm2. This could be attributable to certain parasitic reactions occurring at higher current value. Thus, only a small portion of current intensity was used in degradation/ mineralization of AMT in wastewater. Increase in NaCl concentration increased the amount of active chloro-oxidant species in wastewater, thus increasing the mineralization current efficiency and AMT and TOC removal efficiencies. Furthermore, it decreased the SEC from 3.26 to 0.66 kWh (g TOC removed)1. According to kinetic study, faster degradation rate at higher current density was observed with pseudo-first order rate constant (kf) values, increasing from 0.16 103 to 1.09 103 sec1 with increase in current density from 1.47 to 5.88 mA cm2. Attack of active chlorine on aromatic moiety and subsequent hydroxylation of AMT molecule by active oxygen resulted in formation of major degradation products with lower m/z values, hence lowering the TOC content in synthetic wastewater. Acknowledgements Authors are grateful to University Grants Commission, India for providing research fellowship (MANF-2014-15-SIK-PUN-43596) to the first author. Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.electacta.2018.11.114. References [1] G. Chen, Electrochemical technologies in wastewater treatment, Separ. Purif. Technol. 38 (2004) 11e41. [2] O. Ganzenko, D. Huguenot, E.D. Hullebusch, G. Esposito, M.A. Oturan, Electrochemical advanced oxidation and biological processes for wastewater treatment: a review of the combined approaches, Environ. Sci. Pollut. Control Ser. 21 (2014) 8493e8524. [3] M.A. Oturan, E. Brillas, Electrochemical advanced oxidation processes (EAOPs) for environmental applications, Port. Electrochim. Acta 25 (2007) 1e18. [4] I. Sires, E. Brillas, M.A. Oturan, M.A. Rodrigo, M. Panizza, Electrochemical advanced oxidation processes: today and tomorrow. A review, Environ. Sci. Pollut. Control Ser. 21 (2014) 8336e8367. [5] M. Panizza, G. Cerisola, Direct and mediated anodic oxidation of organic pollutants, Chem. Rev. 109 (2009) 6541e6569.
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Electrochimica Acta journal homepage: www.elsevier.com/locate/electacta
Amoxicillin electro-catalytic oxidation using Ti/RuO2 anode: Mechanism, oxidation products and degradation pathway Ravneet Kaur, Jai Prakash Kushwaha*, Neetu Singh Chemical Engineering Department, Thapar Institute of Engineering and Technology, Patiala, Punjab, India
a r t i c l e i n f o
a b s t r a c t
Article history: Received 25 September 2018 Accepted 17 November 2018 Available online 19 November 2018
Present work investigates the application of electro-catalytic oxidation (EO) technique using dimensionally stable anode, titanium coated with ruthenium dioxide (Ti/RuO2), for abatement of amoxicillin trihydrate (AMT), a much commonly prescribed antibiotic detected in water and wastewater. AMT removal efficiency (%ARE) and TOC (Total organic carbon) removal efficiency (%TRE) were measured by varying process parameters such as initial pH, current density (i), initial AMT concentration (C0) and supporting electrolyte (NaCl) concentration (S0). Mineralization current efficiency (%MCE) and specific energy consumption (SEC) values were evaluated and compared for different values of i and S0. Furthermore, decay kinetics of AMT was studied by varying i and C0. Moreover, AMT degradation and mineralization mechanism was explored in detail. Additionally, a possible pathway of AMT degradation/ mineralization was proposed by identifying the intermediates formed during EO reactions using UPLC-QTOF-MS. Besides, economic feasibility of EO treatment method was analysed by calculating the operating cost. The optimum current density and initial pH were found to be 5.88 mA cm2 and 7.0, at which, 60% ARE and 48% TRE were achieved in 60 and 240 min of electrolysis, respectively. Mineralization current efficiency was observed decreasing from 11.77% to 7.67% with increasing i value. © 2018 Elsevier Ltd. All rights reserved.
Keywords: Electro-catalytic oxidation Amoxicillin trihydrate Ti/RuO2 Mineralization current efficiency Specific energy consumption Degradation pathway
1. Introduction Water pollution is one of the major concerns of mankind in the present era. Urbanization, industrialization and population growth has led to consumption of enormous gallons of water, which in turn leads to generation of hefty volumes of wastewater. A number of treatment technologies have been developed and implemented at production sites to eliminate detrimental pollutants from wastewater, in order to make it suitable to be discharged in to aquatic environments. Since the introduction of electrochemical advanced oxidation processes (EAOPs) in the late 1970s, they have fetched much of the attention of scientific groups worldwide for treatment of bio-refractory and toxic organic pollutants from water [1e4]. EAOPs are based on in situ generation of hydroxyl radical ( OH), a powerful oxidizing agent which unselectively destroys the organic content in wastewaters. Among different EAOPs, electro-oxidation (EO) method of water treatment has become quite popular for its application in industrial wastewater treatment and target pollutant
* Corresponding author. E-mail addresses: [email protected] (J.P. Kushwaha), neetu.singh1479@gmail. com (N. Singh). https://doi.org/10.1016/j.electacta.2018.11.114 0013-4686/© 2018 Elsevier Ltd. All rights reserved.
destruction [5,6]. Major advantages of EO are: 1) no sludge generation, hence environmentally compatible; 2) minimum or no addition of chemicals required; 3) versatile; 4) amenable to automation; 5) operates at mild conditions and simple equipment, hence safe; 6) robust as reaction can be easily terminated by switching off the power [7]. EO technique destroys organic pollutants by two different pathways, i.e., “direct oxidation” at the anode surface and “indirect oxidation” in the bulk solution by oxidizing agents generated on the anode surface [5,8]. Performance and efficiency of EO technique have been investigated for degradation of persistent organic pollutants, such as pharmaceutical compounds, from wastewater [9e15]. Much focus has been given on removal of antibiotics by EO [16e20], since the studies in different parts of globe have reported presence of antibiotics in surface water, ground water and effluent from wastewater treatment plants (WWTPs) in the range of few ng L1 to 100 mg L1 [21e27]. Conventional physico-chemical and biological treatment methods used in WWTPs are unable to remove antibiotics from wastewater because of their biorefractory nature, thus resulting in their accumulation in surface and ground water [21,28]. It is a problematic issue as research has revealed continuous rise in antibiotic-resistant bacteria, which is dangerous for both humans
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and animals, because at present, there are quite limited alternatives to available antibiotics [29e31]. Amoxicillin trihydrate (AMT) is a semi-synthetic broad spectrum b-lactam antibiotic. It is one of the most prescribed antibiotics for children, adults and animals, for a number of bacterial infections, such as pneumonia, middle ear infections, strep throat, skin infections and urinary tract infections. It is commonly detected in sewage treatment plant effluents and surface waters [22,32], as only 20e30% of AMT is metabolized in human and livestock body systems and rest is excreted [33]. There were very few studies available in literature dealing with AMT removal using electrochemical oxidation technique [34e37]. In an interesting study, Sopaj et al., compared the capability of different anode materials such as BDD (Boron doped diamond), Ti/ RuO2-IrO2 (DSA), Pt, PbO2, carbon-fiber, carbon-graphite and carbon-felt for AMT removal. Complete mineralization was attained using BDD in 6 h after applying 41.66 mA cm2 of current density. Nonetheless, for same reaction conditions, less than 25% mineralization was reported for DSA. Similarly, Ganiyu et al., compared the performance of BDD, Pt, DSA and Ti4O7 as anode materials, and carbon-felt as cathode material for EO treatment of AMT. Carbonfelt cathode assisted in formation of H2O2 from compressed air circulated in to the EO cell. Ti4O7 exhibited better results than Pt and DSA, but was found to be less effective than BDD. Frontistis et al., performed AMT oxidation for current density values of 30 and 50 mA cm2 by means of BDD anode, and proposed possible pathways of degradation of AMT by BDD. Another study on BDD application for AMT removal was reported by Panizza et al. [60] wherein complete mineralization by electro-Fenton process was reported. However, the requirement of highly acidic pH conditions (pH z 3.0) for Fenton's reaction is a major drawback of this process. Likewise, BDD anodes have proven their competence by being dimensionally stable and giving supreme performance for destruction of other varieties of antibiotics as well [38e42]. However, these are comparatively much more expensive, which restricts their usage and questions their application on large-scale [5]. In another remarkable study, Santos et al., prepared a dimensionally stable anode (DSA), Ti/Pt/SnO2-Sb2O4 and reported successful removal of AMT from wastewater. Nevertheless, ~50% TOC abatement was achieved after 24 h of EO at 30 mA cm2of current density. More recently, titanium coated with ruthenium dioxide (Ti/ RuO2) has emerged as a promising anode material for degradation of organic pollutants because of its low cost, dimensional stability, high efficiency in strong acidic conditions, high electro-catalytic activity for generation of chloro-oxidant species (Cl2, OCl, HOCl) and other strong oxidizing agents ( OH, H2O2, O3) [43e46]. Such strong mechanical and chemical properties along with high oxygen evolution overpotential (z1.9 V) and low cost, make Ti/RuO2 a better alternative to rest of the anode materials. The degradation of organic pollutant (R) by EO treatment technique using Ti/RuO2 electrodes can follow the route of direct oxidation at the anode surface and/or indirect oxidation in the bulk solution. Direct anodic oxidation takes place via direct electron transfer from anode to organic molecule (R) as shown in following reaction: Ti/RuO2 þ R / Ti/RuO2.(R)ad / nCO2 þ oxidation products þ yHþ þ ze
857
chemisorbed OH, it gets oxidised to form RO, which is also called electrochemical “conversion” (Eq. (5)) [47]. Further oxidation of RO would lead to formation of CO2, H2O and other degradation products and ions. Ti/RuO2 þ H2O / Ti/RuO2.( OH)ad þ Hþ þ e (Physisorption)
(2)
Ti/RuO2.( OH)ad / Ti/RuO3 þ Hþ þ e (Chemisorption)
(3)
Ti/RuO2.( OH)ad þ R / Ti/RuO2 þ nCO2 þ oxidation products þ yHþ þ ze
(4)
Ti/RuO3 þ R / Ti/RuO2 þ RO
(5)
However, the inevitable but undesirable reaction of oxygen evolution occurs simultaneously as shown in following reactions: Ti/RuO2.( OH)ad / Ti/RuO2 þ 1/2O2 þ Hþ þ e
(6)
Ti/RuO2þ1 / Ti/RuO2 þ 1/2O2 þ Hþ þ e
(7)
Despite the advantages, studies on application of Ti/RuO2 for removal of antibiotics from wastewater are very few [45,49]. The novelty of present study lies in application of Ti/RuO2 electrodes for degradation and mineralization of AMT by EO treatment methodology and generation of plausible degradation pathway using Ti/ RuO2 anodes. Effect of important operating parameters, such as current density (i), initial pH of synthetic wastewater, supporting electrolyte concentration (S0) and initial AMT concentration (C0) on AMT removal efficiency (%ARE) and TOC removal efficiency (%TRE) of Ti/RuO2 anode was studied in detail. AMT decay kinetics was studied by varying i (mA cm2) and C0 (mg L1). Besides, mineralization current efficiency (%MCE) and specific energy consumption (SEC) were evaluated. Furthermore, the economic feasibility of the EO treatment technique using Ti/RuO2 anode was checked by calculating the operating cost. Major transformation products were identified using UPLC-Q-TOF-MS and a plausible degradation pathway of AMT was proposed. 2. Experimental 2.1. Materials Fresh synthetic wastewater was prepared at room temperature for each electro-oxidation experiment. Predefined quantity of AMT antibiotic (DSM Sinochem Pharmaceuticals, Punjab, India) and NaCl (Loba Chemie Pvt. Ltd., Mumbai, India) were dissolved in 1.5 L of double distilled water by placing the mixture over a magnetic stirrer for 1 h. NaCl was added as supporting electrolyte to enhance the conductivity of the synthetic wastewater. Initial pH of wastewater was adjusted using 0.1 M HCl or 0.1 M NaOH. All the chemicals used in the procedure were of analytical grade. Dimensionally stable Ti/RuO2 electrodes were purchased from Titanium Tantalum Products Ltd. Company, Chennai, India. 2.2. Electrochemical treatment
(1)
Electrolysis of water produces active oxygen in the form of hydroxyl radical at the anode surface which can either get physisorbed on to anode surface (Eq. (2)) or can get chemisorbed into the oxide matrix of the Ti/RuO2 anode, thus resulting in formation of Ti/RuO3 (Eq. (3)). When the pollutant species (R) come in contact of the physisorbed active oxygen, it results in to complete “combustion” as shown in Eq. (4). However, when R reacts with
The electrooxidation experiments were performed in a cuboid shaped reactor (13 cm 13 cm 13 cm) made from acrylic sheet. Set of four 1.5 mm thick Ti/RuO2 rectangular electrodes (10 cm 8.5 cm) were connected in parallel in bipolar mode such that two alternate electrodes worked as anode and the other two as cathode. The inter-electrode gap was fixed as 1 cm. 1.5 L of freshly prepared synthetic wastewater was introduced into the reactor and continuously mixed throughout the experiment with the help of a
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magnetic stirrer. Electrodes were immersed in the wastewater such that the bottom of the reactor was 2.5 cm apart from lower edges of the electrodes, which ensured proper rotation of the magnetic bead. Current controlled conditions for electrolysis were maintained using a regulated D.C. power supply (DIGITECH, Roorkee, India, Model: 4818A10; 0e20 V, 0e5 A). Electrooxidation reactions began when the power supply was switched on. The set of electrodes was properly washed before performing each experiment. The schematic diagram the electrochemical set-up is shown somewhere else [45]. In order to study the degradation of AMT using Ti/RuO2 electrodes, a set of experiments was performed over a range of parameters initial pH ¼ 2.0e9.0, current density (i) ¼ 1.47e5.88 mA cm2 (corresponding to applied current intensity in the range of 0.25e1.00 A), initial AMT concentration (C0) ¼ 10e50 mg L1 and NaCl concentration (S0) ¼ 1.0e2.0 g L1. Residual AMT concentration and TOC content were analysed by withdrawing samples from the reactor at pre-set time intervals. The mineralization reaction of AMT involves its conversion to þ CO2 and inorganic ions such as SO2 4 and NH4 as shown in following equation: þ 2 C16H19N3O5S$3H2O þ 28H2O / 16CO2 þ 3NHþ 4 þ SO4 þ 69H þ 70e (8)
Mineralization current efficiency (MCE, in %), which is defined as ratio of charge used for mineralization of a pollutant to total charge passed during the electrolysis process, was calculated using Eq. (9).
MCE ð%Þ ¼
TOC 0 TOC f zFV R 4:32 107 nIt
100
(9)
where, TOC0 and TOCf are initial and final TOC concentration (mg L1), respectively, F is the Faraday's constant (96,485 C mol1), z is the number of electrons involved in oxidation of AMT (z ¼ 70 (Eq. (8)), VR is volume of the reactor (L), n is number of carbon atoms in the AMT molecule (n ¼ 16), I is the applied current intensity (A) and Dt is the electrolysis time (h). In order to homogenize units, a conversion factor of 4.32 107 (3600 s h1 12,000 mg mol1) was used in the above equation. The economic feasibility of the electrooxidation process using Ti/RuO2 electrodes was checked by evaluating the specific energy consumption (SEC) using Eq. (10). SEC is defined as the amount of energy consumed in kWh for reduction of 1 g of organic load from wastewater.
SEC kWhðg TOC removedÞ1 ¼
UIt TOC 0 TOC f
VR
(10)
where, U is average voltage of electrochemical cell (V) and t is the electrolysis time (h). The AMT decay by EO process at any instant of time (t) followed pseudo-first order reaction kinetics with rate constant kf (sec1).
2.3. Instrumentation and analytical measurements Initial pH of synthetic wastewater was measured using Thermo Scientific Orion 5 star pH meter. The residual AMT concentration was determined using UVevis spectrophotometer (Perkin Elmer, Lambda 35). A calibration curve for AMT was prepared at lmax ¼ 228 nm. %AMT removal efficiency (ARE) was calculated using the following equation:
ARE ð%Þ ¼
ðC 0 C t Þ 100 C0
(11)
where, C0 and Ct are initial and final AMT concentration (mg L1) respectively. The amount of Total organic carbon (TOC) in the solutions was calculated using TOC-VCPH, PC-controlled TOC Analyser, Shimadzu (oxidation performed at 680 C through platinum catalysed combustion). Mineralization degree of AMT wastewater was evaluated by examining %TOC removal efficiency (TRE) over a period of time by following relationship:
TRE ð%Þ ¼
TOC 0 TOC f TOC 0
100
(12)
AMT transformation products were analysed by ultraperformance liquid chromatography coupled with mass spectrometry, using UPLC-XEVO-G2-XS/QTOF-MS (Waters) equipped with BEH C18 column (100 mm 2.1 mm; 1.7 mm). The mobile phase contained 0.1% of formic acid in water (A) and acetonitrile (B). The flow rate was set as 0.3 mL min1. The initial chromatography method condition held the mobile phase (90% A and 10% B) constant for 2.5 min. Thereafter a linear gradient to 50% B up to 5.5 min was followed by a linear increase to 90% B, kept constant for next 2.5 min. Finally, the mobile phase was returned to initial conditions of 10% B for 2 min. The MS analyser was operated in positive ion mode across the range 100e1000 m/z with source and desolvation temperature set as 120 C and 350 C, respectively. Nitrogen was employed as a carrier gas. The capillary voltage and cone voltage were set as 2000 V and 30 V, respectively. Flow rate for cone gas and desolvation gas were 60 L h1 and 800 L h1, respectively. Fourier transform infrared (FT-IR) spectral study was conducted using Spectrum 2 (Perkin-Elmer) FT-IR spectrophotometer. The transmission spectra obtained was plotted against the wavenumber in the range of 4000e400 cm1. 3. Results and discussion 3.1. Effect of initial pH on degradation and mineralization of AMT Efficiency of electrooxidation process for abatement of organic pollutants is highly pH dependent. The nature of various oxidation species generated on the anode surface and in the bulk solution strongly depends on the pH of wastewater under treatment. Effect of initial pH (2e9) of the solution on %ARE and %TRE with time was studied at current density, i ¼ 5.88 mA cm2, S0 ¼ 2 g L1 and C0 ¼ 50 mg L1, and is shown in Fig. 1a and b, respectively. As shown in Fig. 1a, %ARE increased with an increase in pH of the solution from 2 to 7. After 120 min of EO process, 33.22% of AMT was degraded at pH ¼ 2, whereas, at pH ¼ 7, similar removal was achieved in just 10 min of EO. Almost 60% of AMT was degraded in 45 min of EO at pH ¼ 7. Thereafter no significance increase in %ARE was observed. 40.77%, 48.89% and 52.59% AMT was degraded in 120 min of electrolysis for pH values of 3, 4 and 5.5 (natural pH of synthetic wastewater), respectively. With further rise in pH (8 and 9), there was no improvement in removal efficiency. Similar trend was observed in case of %TRE values (Fig. 1b). In case of highly acidic medium (pH ¼ 2), 11.53% of TOC was removed in 120 min of EO. However, for the similar electrolysis time and conditions, a significant enhancement in %TRE was observed when pH was raised up to neutral level 7, as 41.21% of TOC content in synthetic wastewater got degraded. For pH values of 3, 4 and 5.5, 18.07%, 27.29% and 37.53% of TRE was achieved in 120 min, respectively. ~40% TRE was attained for pH values of 8 and 9. Therefore, no significant change in TRE was observed when pH was raised above 7.0.
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following reactions: Ti/RuO2 þ HOCl / Ti/RuO2.( OCl)ad þ Hþ þ e
(17)
Ti/RuO2.( OCl)ad / Ti/RuO3 þ Cl
(18)
The dominance of these species is highly pH dependent. Cl2 is predominant in highly acidic conditions (pH < 1), whereas, HOCl dominates in the range of 3 < pH < 5 and ClO in pH > 9 [4,47,53]. Oxidation of AMT in acidic conditions mainly occurs through Cl2, which has low oxidation potential, and chemisorbed hydroxyl radical, which are predominant in acidic conditions. Furthermore, dissolved chlorine gets reduced in acidic conditions, resulting in oxygen evolution in the solution (Eq. (19)), which resulted in low removal efficiency of EO system for AMT and TOC content [50]. 2Cl2 þ 2H2O / O2 þ 4HCl
(19)
Chemisorbed hydroxyl radicals lead to partial conversion of AMT molecules (Eq. (5)), hence resulting in quite low value of %TRE as compared to %ARE. After the pH is raised, hypochlorous acid becomes the major chloro-oxidant responsible for degradation of AMT in synthetic wastewater, whose oxidation potential is higher than that of Cl2. Thus, a significant increase in values of %ARE and % TRE was observed with increase in pH values up to 7. Moreover, in this pH range, physisorbed active oxygen and physisorbed oxychloro species play their part in AMT degradation and mineraliza tion (Eqs. (4), (17) and (18)). The increase in amount of OH radicals with increase in pH is also a key factor behind improved values of % ARE and %TRE [54]. As discussed above, %ARE and %TRE increased with the increase in initial pH of the wastewater up to pH 7.0, with no significant difference for higher pH values of 8 and 9. Thus, the optimum pH value for the present study was found to be 7 and further experiments were conducted at neutral pH only. 3.2. Effect of current density (i) on degradation and mineralization of AMT
Fig. 1. Effect of initial pH on (a) AMT removal efficiency, %ARE, and (b) TOC removal efficiency, %TRE (Other parameters: i ¼ 5.88 mA cm2, S0 ¼ 2 g L1, C0 ¼ 50 mg L1).
Ti/RuO2 electrodes exhibit high electrocatalytic activity for generation of reactive chlorine species which play an important role in indirect oxidation of organic pollutants [5,8,50]. Chlorine is electro-generated by direct oxidation of Cl ion on the anode surface as demonstrated in reactions (13) and (14) [50e52]. Reaction of chlorine with water results in formation of hypochlorous acid (HClO) in the bulk solution (Eq. (15)). Furthermore, deprotonation of HOCl generates hypochlorite ion (ClO) (Eq. (16)) Ti/RuO2 þ Cl / Ti/RuO2.( Cl)ad þ e
(13)
2Ti/RuO2.( Cl)ad / 2Ti/RuO2 þ Cl2
(14)
Cl2 þ H2O / HOCl þ Hþ þ Cl
(15)
HOCl þ OH / ClO⁻ þ H2O
(16)
The oxygen transfer to molecules of pollutant species can also take place through adsorbed oxy-chloro radicals as shown in
The effect of applied current density, i (1.47e5.88 mA cm2), on EO process parameters by Ti/RuO2 electrodes at pH ¼ 7, C0 ¼ 50 mg L1, S0 ¼ 2 g L1, was studied systematically. Fig. 2a shows the effect of different values of applied i (mA cm2) on %ARE values with electrolysis Significant increase in %ARE values was observed when i values were increased from 1.47 to 5.88 mA cm2. AMT removal improved from 9.77% at 30 min to 22.75% at 120 min for i ¼ 1.47 mA cm2. When EO was carried on further, 26.32% of AMT content got degraded in 150 min, with no further increase up to 180 min. Similarly, for i ¼ 2.94 mA cm2, equilibrium was attained in 150 min of EO with 38.44% ARE. However, for i values of 4.41 mA cm2 and 5.88 mA cm2, equilibrium stage was achieved in 75 min and 45 min, with 53.65% and 61% AMT removal, respectively. AMT removal rate got enhanced with the increase in applied current because the amount of oxidant species (chemisorbed and physisorbed OH, Cl2, HOCl, ClO) generated at the anode surface according to Eqs. (2) and (3), (13)e(18), increased substantially when applied current intensity was increased. However, not more than ~61% of ARE was attained for 5.88 mA cm2 of applied current density for present experimental conditions. This might be because of the inevitable reaction of oxygen evolution taking place at anode surface at higher values of current intensity (Eqs. (6) and (7)). Efficiency of EO for TOC abatement was tested in the same range of current density (1.47e5.88 mA cm2) for prolonged electrolysis time up to 240 min (Fig. 2b). 19.15%, 36.46%, 46.18% and 48.62% TRE was achieved for current densities of 1.47, 2.94, 4.41 and
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Fig. 2. Effect of applied current density, i (mA cm2) on (a) AMT removal efficiency, %ARE, (b) TOC removal efficiency, %TRE, and (c) Mineralization current efficiency, %MCE and Specific energy consumption, SEC (kWh (g TOC removed)1) (Other parameters: pH ¼ 7.0, S0 ¼ 2 g L1, C0 ¼ 50 mg L1).
5.88 mA cm2, respectively in 240 min of EO. Threefold escalation in %TRE was observed, when i was increased from 1.47 to 5.88 mA cm2. In neutral pH range, physisorbed active oxygen and HOCl play the key role in mineralization of organic species. The amount of these oxidants improved significantly with increase in applied current, hence resulting in substantial improvement in % TRE values. Comparative analysis of %MCE and SEC for different values of i after 120 min of EO is shown in Fig. 2c. For different values of current density, SEC increased from 0.228 to 0.558 kWh (g TOC removed)1 with an increase in i from 1.47 to 5.88 mA cm2. However, %MCE decreased from 11.77% to 7.67% for the same values of current density. Even though, TOC removal rate was enhanced with the increase in applied current density, still the energy consumption of EO system increased because of elevation in average EO cell potential, U from 3.1 V to 5.2 V, with increase in i from 1.47 to 5.88 mA cm2. As discussed above, the unescapable reaction of oxygen evolution (Eqs. (6) and (7)) is enhanced at higher current intensities which competitively consumes the extra electrons. Although, at higher i, the rate of generation of OH radicals is elevated, the unreacted radicals tend to form hydrogen peroxide and hydroperoxyl radical according to Eqs. (20) and (21) [55], whose oxidation power is much less as compared to that of hydroxyl radical [56]. 2 OH / H2O2
(20)
H2O2 þ OH / HO2 þ H2O
(21)
Aggravation of such parasitic reactions at higher current intensity causes the loss in mineralization efficiency of EO process, thus increasing the overall cost of treatment. 3.3. Effect of supporting electrolyte concentration For the present study, NaCl was employed as a supporting electrolyte for all EO experiments. It served two purposes: (1) enhancement of electrical conductivity of the synthetic wastewater to perform the electrolysis; (2) deliver active chlorine species as oxidizing agents for abatement of organic pollutant (AMT) present in synthetic wastewater. In order to study its effect on performance of EO process, a range of 0.5e2 g L1 of NaCl concentration (S0) in synthetic wastewater was selected. Other operating parameters were set as: i ¼ 5.88 mA cm2, pH ¼ 7, C0 ¼ 50 mg L1. As depicted in Fig. 3a and b, the increase in S0 resulted in increase in both %ARE and %TRE values. ARE value got enhanced from 52.51% to 60.88% after 120 min of electrolysis, when S0 was increased from 0.5 to 2 g L1 (Fig. 3a). For the same set of S0 values and electrolysis time, % TRE values also showed threefold improvement from 16.31% to 48.67% (Fig. 3b). The mediated oxidation of AMT was carried out by the oxy-chloro species generated at the Ti/RuO2 surface according to Eqs. (13)-(18). It was evident that the amount of these reactive chloro oxidants increased with the increase in concentration of NaCl in synthetic wastewater, thus enhancing the efficiency of EO for removal of AMT and TOC. Correlation of S0, %MCE and SEC is shown in Fig. 3c. %MCE increased with exponential decay in SEC (kWh (g TOC removed)1), when S0 (g L1) was increased. With increase in amount of NaCl, the
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Fig. 3. Effect of NaCl concentration, S0 (g L1) on (a) AMT removal efficiency, %ARE, (b) TOC removal efficiency, %TRE, and (c) Mineralization current efficiency, %MCE and Specific energy consumption, SEC (kWh (g TOC removed)1) (Other parameters: i ¼ 5.88 mA cm2, pH ¼ 7.0, C0 ¼ 50 mg L1).
conductivity of synthetic wastewater got significantly increased from 1.32 mS cm1 (S0 ¼ 0.5 g L1) to 4.08 mS cm1 (S0 ¼ 2 g L1), thus lowering the average cell potential (U) from 10.2 V (S0 ¼ 0.5 g L1) to 5.2 V (S0 ¼ 2 g L1), which in turn decreased the amount of energy consumed by the electrolytic system. Higher quantity of NaCl amplified the generation rate of chloro-oxidant species in wastewater, consequently increasing the TOC decay rate, and thus the mineralization efficiency of EO by Ti/RuO2 anodes. Therefore, the overall cost of EO treatment method decreased and the %MCE increased, when the amount of NaCl in synthetic wastewater was increased from 0.5 g L1 to 2 g L1. 3.4. Influence of initial AMT concentration and decay kinetics Degradation of AMT antibiotic by EO treatment technique at any instant of time (t) can be demonstrated by pseudo-first-order kinetic model given as below:
ln
C0 ¼ kt Ct
(22)
For any value of C0 and i, concentration of AMT in the wastewater first decreases with time and then attains an equilibrium value of Ce, leaving behind a fraction of initial AMT concentration, which could not be removed at experimental conditions. Therefore,
a more appropriate way for representing the above Eq. (22) is the modified form given below [57]:
C Rt ¼ C Re 1 ekf t
(23)
where, CRt and CRe (mg L1) are the amount of AMT removed at any time, t and at equilibrium conditions, respectively, and kf is pseudofirst order rate constant. Non-linear regression using Marquardt's percent standard deviation (MPSD) error function was performed by fitting the experimental data in following equation [58]:
vffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi !2ffi u n u 1 X C C Rt;i;exp Rt;i;cal MPSD ¼ 100t n p i¼1 CRt;i;exp
(24)
where, ‘n’ and ‘p’ are number of measurements and number of parameters in the model, respectively, and the subscripts ‘exp’ and ‘cal’ represent the experimental and calculated values of CRt. Kinetic study for degradation of AMT was performed by varying the initial concentration of AMT (C0) in the range of 10e50 mg L1 at i ¼ 5.88 mA cm2, pH ¼ 7.0, S0 ¼ 2 g L1 and also by varying current density in the range of 1.47e5.88 mA cm2 at same experimental conditions with C0 ¼ 50 mg L1. The kinetic data fitted to pseudo-first-order kinetic model by varying C0 and i are shown by
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solid line in Fig. 4a and b, respectively. The best fit values of rate constant, correlation coefficient and MPSD are shown in Table 1. The equilibrium condition was achieved in 30 min for C0 ¼ 10 mg L1, 60 min for C0 ¼ 20, 30 and 40 mg L1, and 90 min for C0 ¼ 50 mg L1. Hence, the time range for the kinetic study of AMT degradation was set accordingly as presented in Fig. 4a. As evident from data in Table 1, R2 values (~0.99) and low MPSD values justify the aptness of pseudo-first-order kinetic model for AMT degradation by EO using Ti/RuO2 electrodes. Furthermore, the proximity of CRt,exp and CRt,cal values reinforces the fitness of pseudo-first-order kinetic model for current study. The values of kf decreased from 2.36 103 to 1.09 103 sec1, with the increase in C0 from 10 to 50 mg L1 (Table 1). These results suggest that the degradation rate of AMT by EO at its lower concentration levels was faster than its higher concentration levels. The nature and amount of oxidation species ( OH, HOCl, ClO, H2O2, HO2, etc.) generated at the anode surface and in the bulk solution remains the same when a fixed current is supplied to the EO system. When substrate (AMT) concentration in wastewater is low, the
Table 1 Pseudo-first order kinetic parameters. Parameter
Value
kf 103 (sec1)
R2
MPSD
i (mA cm2)
1.47 2.94 4.41 5.88 10 20 30 40 50
0.16 0.17 0.63 1.09 2.36 1.98 1.29 1.17 1.09
0.997 0.997 0.998 0.996 0.997 0.994 0.999 0.992 0.996
9.28 12.17 6.46 6.88 3.56 10.53 4.34 5.04 6.88
C0 (mg L1)
oxidants attacking its molecules for degradation are in plenty, thus resulting in faster removal. Conversely, the ratio of substrate concentration and oxidants decreases for higher values of AMT concentration, causing the slower degradation rate, and thus lower kf values. When i was increased from 1.47 to 5.88 mA cm2, the value of kf also increased from 0.16 103 to 1.09 103 sec1 (Table 1). As discussed in section 3.2, the increase in value of applied current intensity amplifies the generation of oxidants ( OH, HOCl, ClO, H2O2, HO2, etc.) responsible for degradation of AMT molecules in synthetic wastewater. This in turn leads to faster degradation of AMT, and hence, higher values of pseudo-first-order rate constant.
3.5. Plausible AMT degradation pathway
Fig. 4. Kinetics of AMT degradation (a) Effect of initial AMT concentration, C0 (mg L1) (b) Effect of current density, i (mA cm2).
In order to derive the degradation scheme of AMT during EO process by Ti/RuO2 electrodes, synthetic wastewater with 200 mg L1 of AMT was taken in reactor for electrolysis at neutral pH, 5.88 mA cm2 of current density and 2 g L1 of NaCl. Samples after electrolysis reaction were collected for analysis by UPLC-QTOF-MS. The mass spectrum of initial sample before reaction (Fig. SM e 1a) shows peaks corresponding to AMT molecule and its fragmentation ions. The mass/charge ratio (m/z) of 366 corresponds to the protonated molecule of AMT [M þ H]þ. Additional peak at m/ z value of 349 represents the fragmentation ion AF 1 [M e NH3 þ H]þ of AMT molecule. The opening of b-lactam ring of AMT leads to formation of other two fragments AF 2 and AF 3, corresponding to m/z values of 208 and 160, respectively (Table 2). Besides, m/z value of 388 is attributable to the sodium adduct ion [M þ Na]þ. The retention time of AMT was 0.87 min (Fig. SM e 1b). The UPLC chromatogram of sample withdrawn after 15 min of EO reaction is shown in Fig. SM e 2. Three major AMT degradation products (ADPs) were identified, whose retention time, elemental composition and proposed structure is shown in Table 2. The plausible degradation/mineralization pathway of AMT by EO using Ti/RuO2 electrodes is depicted in Fig. 6. The reactions occurring during EO process are quite complex and involve a number of oxidant species, such as OH, HOCl and Cl2, which degrade the organic pollutants via direct or mediated oxidation. Based on the information gathered from UPLC-Q-TOFMS analysis of samples, a possible reaction pathway for AMT degradation was proposed. As shown in Fig. 6, loss of NH3 group, and attack of active chlorine leading to decarboxylation of AMT molecule, resulted in formation of a reaction intermediate ADP 1 (m/z ¼ 339). Subsequent hydroxylation of ADP 1 by active oxygen in the form of hydroxyl radicals, generated another degradation product, ADP 3 with m/z ¼ 325. The replacement of entire aromatic moiety of AMT molecule by active chlorine formed ADP 2 (m/ z ¼ 269). Alternative route for formation of ADP 2 could also be via direct attack of Cl ions on the AMT molecule, thus leading to decarboxylation, loss of aromatic moiety and amine group. Further
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Table 2 Major AMT degradation products (ADPs) identified by UPLC-Q-TOF-MS. Compound
Rt (min)
m/z [MþH]þ
Elemental composition
ADP 1
0.703
339
C15H15ClN2O3S
ADP 2
0.754
269
C8H10Cl2N2O2S
ADP 3
1.18
325
C13H12N2O6S
AMT
0.87
366
C16H19N3O5S
AF1
e
349
C16H16N2O5S
AF2
e
208
C10H11N2O3
AF3
e
160
C6H9NO2S
attack of oxidation species on degradation products would lead to þ formation of aliphatic acids and inorganic ions (SO 4 and NH4 ). Finally, mineralization occurred after oxidation of aliphatic acids to form CO2 and H2O. Mineralization efficiency of nearly 48% was achieved at optimum conditions of EO in 4 h of reaction, signifying the amount of carbon content in synthetic wastewater which got converted to CO2 (inorganic carbon). Degradation pathway of amoxicillin has also been proposed by Refs. [34,35] by employing BDD and Ti4O7 as anodes for electrooxidation treatment, respectively. Hydroxyl radical physisorbed on to the surface of these nonactive anodes played the key role in degradation and mineralization of the antibiotic in both the cases. However, in present study the role of chloro-oxidants was inevitable. Additionally, FT-IR spectral study of initial and final sample was also performed to analyse the effect of EO reactions (Fig. 5). The stretch extending from 3700 to 2750 cm1 is due to eOeH group and vibrations due to intermolecular hydrogen bonding. The peak at 3309.21 cm1 is due to amide bond (eNeH group) in AMT molecule [43,49]. An additional peak at 1636.24 cm1 is the characteristic absorption peak of eC]O bond of eCOOH group. The
Proposed Structure
further peak at 645.42 cm1 is part of the fingerprint region and is mainly due to carbonyl group and bending vibrations within the molecule. Identical peaks could be noticed in the FT-IR spectrum of final sample withdrawn after EO of synthetic wastewater comprising of 50 mg L1 of AMT and 2 g L1 of NaCl at applied current of 1 A and pH 7 after 120 min. However, the intensities of these peaks were significantly decreased as compared to those in spectrum of initial sample. The spectral peaks thus support the findings discussed in previous section which proved the conversion and combustion of AMT molecules via direct and mediated electrolysis by Ti/RuO2 electrodes using EO methodology. 3.6. Operating cost analysis Economic feasibility of any treatment technique is essential for its applicability at high scale. For a WWTP based on EO technique, operational costs would comprise of cost of electrical energy consumed, cost of electrodes, pumping and mixing cost, wages, etc. Amongst these, the cost of anode material and electrical energy are of much importance. Hence, to calculate the operating cost of EO
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Fig. 5. FT-IR spectra of synthetic wastewater before and after EO treatment (i ¼ 5.88 mA cm2, pH ¼ 7.0, S0 ¼ 2 g L1, C0 ¼ 50 mg L1).
Fig. 6. Proposed degradation pathway of AMT by EO treatment using Ti/RuO2 anodes.
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treatment of wastewater under present study, the total cost of electrical power consumed and that of anode material was calculated. At optimum conditions for EO treatment of AMT (i ¼ 5.88 mA cm2, pH ¼ 7.0, S0 ¼ 2 g L1), 0.404 kWh of electrical energy was consumed for removal of 1 g of TOC from wastewater in 1 h of EO. The average cost of electrical energy in India is INR 5 per kWh. Therefore, the price of electrical energy (CEE) comes out to be INR 2.02 (g of TOC removed)1. Ti/RuO2 electrodes (10 cm 8.5 cm; thickness: 1.5 mm) used in present study were procured for INR 2500 each. Two electrodes were employed as anodes for EO treatment of AMT. According to the manufacturer, the life of these electrodes was 2.5 years at optimum experimental conditions (i ¼ 5.88 mA cm2, pH ¼ 7.0, S0 ¼ 2 g L1). Using EO technique under optimum conditions, ~34% TOC was removed from wastewater comprising AMT in 1 h. The electrode cost (CEL) hence comes out to be INR 28.88 (g of TOC removed)1. Therefore, the total cost of EO treatment (CEE þ CEL) for removal of 1 g of TOC is INR 28.88 (~$ 0.44). 4. Conclusions Maximum removal efficiencies is terms of AMT and TOC were achieved at neutral pH of wastewater. Mineralization current efficiency decreased from 11.77% to 7.67% with increase in current density from 1.47 to 5.88 mA cm2. This could be attributable to certain parasitic reactions occurring at higher current value. Thus, only a small portion of current intensity was used in degradation/ mineralization of AMT in wastewater. Increase in NaCl concentration increased the amount of active chloro-oxidant species in wastewater, thus increasing the mineralization current efficiency and AMT and TOC removal efficiencies. Furthermore, it decreased the SEC from 3.26 to 0.66 kWh (g TOC removed)1. According to kinetic study, faster degradation rate at higher current density was observed with pseudo-first order rate constant (kf) values, increasing from 0.16 103 to 1.09 103 sec1 with increase in current density from 1.47 to 5.88 mA cm2. Attack of active chlorine on aromatic moiety and subsequent hydroxylation of AMT molecule by active oxygen resulted in formation of major degradation products with lower m/z values, hence lowering the TOC content in synthetic wastewater. Acknowledgements Authors are grateful to University Grants Commission, India for providing research fellowship (MANF-2014-15-SIK-PUN-43596) to the first author. Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.electacta.2018.11.114. References [1] G. Chen, Electrochemical technologies in wastewater treatment, Separ. Purif. Technol. 38 (2004) 11e41. [2] O. Ganzenko, D. Huguenot, E.D. Hullebusch, G. Esposito, M.A. Oturan, Electrochemical advanced oxidation and biological processes for wastewater treatment: a review of the combined approaches, Environ. Sci. Pollut. Control Ser. 21 (2014) 8493e8524. [3] M.A. Oturan, E. Brillas, Electrochemical advanced oxidation processes (EAOPs) for environmental applications, Port. Electrochim. Acta 25 (2007) 1e18. [4] I. Sires, E. Brillas, M.A. Oturan, M.A. Rodrigo, M. Panizza, Electrochemical advanced oxidation processes: today and tomorrow. A review, Environ. Sci. Pollut. Control Ser. 21 (2014) 8336e8367. [5] M. Panizza, G. Cerisola, Direct and mediated anodic oxidation of organic pollutants, Chem. Rev. 109 (2009) 6541e6569.
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