Chemical Engineering Journal Advances

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Chemical Engineering Journal Advances 8 (2021) 100195

Available online 18 October 2021

Electrochemical degradation of emerging pollutants via laser-induced graphene electrodes

Chidambaram Thamaraiselvan

^a^,^b^,¹

, Dyuti Bandyopadhyay

^a^,¹

, Camilah D. Powell

^a

, Christopher J. Arnusch

^a^,^*

aDepartment of Desalination and Water Treatment, Zuckerberg Institute for Water Research, The Jacob Blaustein Institutes for Desert Research, Ben-Gurion University of the Negev, Sede-Boqer Campus, Midreshet Ben Gurion 8499000, Israel

bInterdisciplinary Centre For Energy Research, Indian Institute of Science, Bangalore 560052, India

A R T I C L E I N F O Keywords:

Laser-induced graphene Advanced oxidation Emerging pollutants Carbamazepine Methylene blue Hydrogen peroxide

A B S T R A C T

Emerging pollutants of concern (e.g., pharmaceuticals) are challenging to remove by conventional methods, persistent in the aquatic system, and can be toxic to mankind and the environment. Carbamazepine (CBZ) is a widely used pharmaceutical for neuropathic diseases and detected in water systems around the world. Similarly, methylene blue (MB), used in pharmaceutical, laboratory and textile industries poses a threat to the environment. Electrochemical water treatment is a promising technology shown to remove emerging pollutants and organic pharmaceuticals through oxidation and reduction pathways. Recently, laser-induced graphene (LIG) electrodes were fabricated on top of polymer materials and showed antifouling properties and effective oxidizing species generation (i.e., H2O2) at 2–3 V. Here we show that LIG-derived electrodes can remove and degrade organic pollutants from water with an applied voltage via reactive oxygen species generation. The surface area of the LIG electrodes, applied voltage, and the concentration of H2O2 were studied to optimize CBZ and MB degradation. We observed that the LIG electrodes could degrade 64% of CBZ after 6 h at 2.5 V and, together with adsorption to the electrode, 82% was removed from solution after 24 h. Partially degraded CBZ accounted for only 3% of total removal and degradation products were identified by ultra-performance liquid chromatography and mass spectrometry. Similarly, a complete color removal of MB could be obtained in 6 h at 2.5 V corresponding to an 80% dye transformation, which was most likely due to chemical oxidation. Since these LIG electrodes can be rapidly formed on flexible polymer substrates at low cost, this technology might lead to novel water treatment devices and strategies presented herein might potentially lead to solutions for removing emerging contaminants in wastewater.

1. Introduction

In more arid regions of the world, potable water scarcity has become the greatest threat to food security, human health, and the ecosystem. It is estimated that about a quarter of the world’s population, nearly 1.4 billion people, live in regions that will face severe water shortages by 2025 [1]. On the other hand, existing freshwater resources are increasingly polluted by different industrial, agricultural, and human activities compounding the problem. A major and sustainable solution to partially address the problem of water scarcity is water reuse through treatment, specifically water desalination or advanced water treatment technologies, which are becoming essential in modern society for

tackling the potable water demand [2]. Unfortunately, however, current water treatment methods increasingly prove to be insufficient to remove emerging contaminants of concern like pharmaceuticals, PPCPS (personal care products), pesticides, surfactant degradants, and commercial dyes considering their frequent detection in surface waters and water treatment effluents around the world [3].

Methylene blue (MB) is a basic aniline dye used extensively in textile industries and some medical applications [4]. However, exposing the environment to excessive amounts of methylene blue cumulated over time proves toxic necessitating its complete removal from aqueous media. Various remediation strategies have been invoked to study MB removal/degradation such as photocatalysis [5] ozonation [6] and

* Corresponding author.

E-mail address: [email protected] (C.J. Arnusch).

1 Authors with equal contribution

Contents lists available at ScienceDirect

journal homepage: www.sciencedirect.com/journal/chemical-engineering-journal-advances

https://doi.org/10.1016/j.ceja.2021.100195

Received 24 May 2021; Received in revised form 9 September 2021; Accepted 30 September 2021

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electrochemical degradation [7,8]. Among these listed, electrochemical degradation is deemed more advantageous in terms of efficiency, electrode recyclability, and cost effectiveness as exemplified by various studies in literature (Table S1). For example, an electrochemical cell consisting of a grade 316 stainless steel cathode and graphite anode effectively decolorized 50 mg L⁻¹and 400 mg L⁻¹MB solutions by 72.51% and 77.94%, respectively, using only 2.055 kWh/m³of energy at an applied voltage of ~15 V [9]. Other examples include electrochemical cells with Co doped Ti/TiO2 nanotube/PbO2 (TiO2–NTs/- Co-PbO2) electrodes in which 100% MB color removal was achieved using ~13 kWh m⁻³at an estimated applied voltage of 4.75 V [7].

Pharmaceutical compounds such as carbamazepine (CBZ) are found in the influent and effluent of sewage water treatment plants in addition to surface waters and appears to be recalcitrant to both conventional and advanced wastewater treatment processes. CBZ is reported to be the most frequently detected emerging pollutant in wastewater treatment plants (WWTPs) and river water in North America and Europe [10–13].

In Spain, CBZ has been detected in groundwater at concentrations up to 610 ng L⁻¹[14]. These types of pollutants can be very mobile and persistent in environmental matrices and can be a source of harmful, toxic and persistent degradation or transformation products [15,16].

Thus, the detection and removal of said pollutants is very essential, which is why many studies have been conducted on the detection and analysis of pharmaceutical pollutants and their degradation products in recent decades [17]. Unfortunately, satisfactory treatment of wastewater containing toxic organic substances is not yet widely available and, therefore, investigating new and effective methods becomes essential [18].

Several techniques have been employed to remove and degrade CBZ from surface and ground water such as ultrasonication, Fenton’s oxidation, ferrosonication, advanced oxidation processes (UV/H₂O₂), biodegradation, and oxidation by chlorine solution [19]. Unfortunately, most of these processes come with disadvantages like the health risks associated with using large amounts of toxic oxidizing agents (e.g., chlorine) or the generation of large, concentrated waste streams, which is especially true of Fenton driven processes. Biological treatments are less effective for treating pharmaceuticals in wastewater and need additional chemical treatments like ozone to sufficiently remove the pollutants [20]. And, methods like nano-filtration or reverse osmosis are not able to completely filter out certain drugs and dyes from wastewater – the small molecular size and neutral charge of the contaminants en- ables them to pass through the membrane [21]. Alternatively, electrochemical oxidation processes continue to be studied for the degradation of organic pollutants like CBZ (Table S1, [22]). For example, recent studies surrounding the electrochemical oxidation of CBZ have focused on increasing system oxidation efficiency (e.g., via in situ O2 generation with a carbon-polytetrafluoroethylene (C/PTFE) cathode, Ti-mesh/PbO2/MEA cathode assembly [23]. Increasing H2O2 and hydroxyl radical production via introducing various catalysts and electrode combinations is another significant area of interest for the electrochemical oxidation of CBZ (e.g., nitrogen doped carbon, FeSO4, or ferrocene functionalized electrochemically reduced graphene oxide (Fc-ErGO) [24–26]). Overall, in addition to being robust, the electrochemical technique offers high removal efficiency, low temperature requirements [27], and is increasingly becoming more popular due to its versatility, automation, energy efficiency, and cost effectiveness – as exemplified in the low operational costs associated with electro-Fenton enhanced dimensionally stable anode for CBZ degradation (i.e., BOD5/COD =0.68 at US$ 1.46 m⁻³in 4 h) [24,28].

However, choice of electrode material plays a major role in the rate of electrochemical oxidation and associated operational costs [29,30].

Electrode materials including boron doped diamond (BDD) anodes [31], mixed metal oxides [32,33], titanium meshes coated with mixed metal oxides [34], carbon based electrodes (e.g., graphene, graphite) [35], and carbon nanotube composites [36,37] are used for electrochemical oxidation and reduction of organic compounds. However, these

electrodes, along with those listed above for MB and CBZ degradation, can come with operational high costs, issues related with chemical leaching, electrode instability and contaminant sludge by products specifically in the case of electro Fenton type reactions. Graphene – an sp²hybridized allotrope of carbon and is the building block of graphite [38,39] – has unparalleled physical, chemical, mechanical and electrical properties and has been widely studied and applied for various fields such as energy storage devices, sensors, drug delivery, environmental technologies, biomedical devices and so on [40–46]. In fact, laser-induced graphene (LIG) can be readily generated on different carbon materials and polymers using a relatively low-cost CO2 laser and this technique has been applied to a wide variety of energy and environmental applications [47–50]. Previously it was shown that LIG could be fabricated on the aromatic polysulfone-class of polymers, which are highly used in membrane filtration [51–53], electronic devices [54], thin film technologies [55,56], fuel cells [57,58] and biomaterials [59, 60]) expanding the technological versatility of LIG.

In this present study we generate conductive carbon electrodes using the highly versatile technique of fabricating LIG on poly(ether sulfone) (PES) to degrade two known persistent pollutants commonly found in water bodies: CBZ and methylene blue (MB) [61,62]. LIG electrodes were fabricated, characterized, and incorporated into an electrochemical cell to investigate the electrochemical degradation of CBZ and MB. To achieve this aim: (1) laser setting were optimized during LIG fabrication, (2) voltages and electrode surface areas – which affect the electrochemical generation of oxidant species – were varied, and (3) H2O2 concentrations were also measured during the electrochemical degradation of each organic compound of interest. Lastly, degradation reaction mechanisms for CBZ and MB decolouration were proposed.

2. Materials and methods 2.1. Materials

PES sheets (~100 µm thickness) were made from commercial polymer pellets (Product number: 39,119,019) obtained from BASF (Ger- many). Dichloromethane (DCM) was obtained from Bio-lab limited Israel. Sodium sulfate anhydrous (Na₂SO₄, 99.6%) was obtained from Carlo Erba reagents group (Israel). Hydrogen peroxide (H2O2, 30%) was obtained from Merck, (Germany). N,N‑diethyl-p-phenylenediamine (DPD) reagent, ferrous sulfate (FeSO4), and CBZ powder (C15H12N2O, MW 236.27 g/mol) (see supplementary information Fig. S1), trifluoro- acetic acid (TFA), MB dye (chloride salt) (see supplementary information Fig. S1) and copper tape were purchased from Sigma-Aldrich (Israel). Deionized (DI) water was obtained from a Milli-Q ultrapure water purification system (Millipore, Billerica, MA, USA). Carbon glue and epoxy glue were obtained from Melrose, USA. SPE syringes with hydrophobic polystyrene-divinylbenzene adsorbent resin (3 mL/200 mg) were used (Chromabond HR-X, Germany). Acetonitrile (HPLC grade) was obtained from J.T. Baker limited (Israel), and HPLC grade methanol (99.9%) from Beith Dekel, (Israel).

2.2. Fabrication of LIG electrodes

PES sheets were fabricated by dissolving 12.5 w/w% of PES pellets in DCM under stirring at 140 rpm for 4 h. The solution was poured into a glass petri-dish (diameter 14 cm), which was covered by aluminum foil and placed in a fume hood for evaporation of the solvent for 10–12 h.

The resulting dry PES sheet was removed from the petri-dish and used for LIG fabrication. The LIG was generated on the PES sheet with a 50 W, 10.6 μm CO2 laser cutting system (Universal VLS 3.50 laser cutter platform) using the following settings: focused laser, image density of 70 pulses per inch (PPI), a scan rate of 25%, and power 2.5%. The PES-LIG was prepared in air and under ambient conditions (1 atm). Two parallel LIG electrode areas of 8 cm²(2 cm x 4 cm) separated by 1 cm from each other were thus fabricated. Copper tape was attached to each LIG

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electrode with carbon glue and epoxy to serve as electrical connections.

2.3. Methylene blue and carbamazepine degradation using an electrochemical cell

The electrochemical degradation of MB and CBZ was tested using an electrochemical cell operating under batch conditions (Fig. 1). Experi- ments were performed in 70 mL of DI water containing 0.05 M Na2SO4

and 5 mg L⁻¹of MB or 2 mg L⁻¹of CBZ with constant magnetic stirring.

The LIG electrodes were incorporated into the electrochemical system shown in Fig. 1. The power source was connected to the LIG electrodes and the applied voltage and current was monitored by the source and also a multimeter attached to the circuit (Fig. S2). LIG electrodes of 8 cm²were partially immersed (7 cm²) into the solution such that the electrical connections did not contact the electrolyte. Parafilm was used to seal the electrochemical cell and the apparatus was protected from light with aluminum foil. During experiments, the power source was turned on (0.0–2.5 V) and samples (~1 mL) were collected from the cell at 0, 2, 4, and 6 h after which the electricity was turned off. The system remained at room temperature under constant stirring with the final sample being collected at 24 h. The absorbance of the MB samples were measured at 656 nm with an UV-spectrophotometer (UV-1800, Shi- madzu; MB detection limit ~ 0.2 mg L⁻¹). The pH of the solution was measured using a pH-meter (PHS-3D pH meter, Sanxin) at 0, 3, 6, and 24 h, and the sample color was visually compared to previously collected MB solution samples.

To test the effect of pH variations on MB and CBZ degradation, solution pH was adjusted to 1 with HCl or to 13.7 with NaOH. 2 mL samples were taken, and the absorbance spectrum was measured as before. To estimate the adsorption of both CBZ and MB on the LIG electrodes, control experiments were performed as described above but in the absence of electricity. Here too, samples were collected at 0, 2, 4, 6 and 24 h and analysed by HPLC (CBZ detection limit ~0.25 mg L⁻¹) and UV-spectrophotometry, respectively. Additional control experiments were performed by adding 2 ppm H₂O₂to the electrolyte without applying any voltage and samples were collected and analysed as noted above. All experiments were conducted in triplicate with the average value presented. Error bars correspond to one standard deviation in each direction for the dependent variable in question.

Lastly, to benchmark the performance of the electrochemical cell, the electrochemical degradation of MB was compared to conventional Fenton’s chemistry under batch conditions. Experiments were

performed in a similar fashion to the electrochemical experiments with the following exceptions: initial H₂O₂concentrations were varied (i.e., 1–3 ppm), electrodes and power sources were omitted, and 0.2 mM of FeCl2 was added into the solution.

2.4. HPLC and UPLC-MS analysis

Sample solutions were desalted using a solid phase extraction cartridge CHROMABOND C18. The cartridge was washed twice with 2.5 mL methanol (UPLC grade) followed by 2.5 mL DI water before analysis.

Each sample (1 mL) was filtered through the cartridge followed by 5 mL of DI water. Compounds were eluted from the cartridge with HPLC grade methanol. Samples were analyzed with HPLC (Agilent 1100 series) using C-18 reverse-phase column (YMC; 250 ×4.6 mm I.D., S-4 μm, 8 nm, Ophir analytical, Israel). Mobile phase solutions used for the chromatography were: 0.08% (v/v) trifluoracetic acid in acetonitrile and 0.1%

(v/v) trifluoracetic acid in distilled water. Injection volume of each sample was 10 µL, column temperature was 30 ^◦C, and each sample was run for 45 min.

For LC-MS analysis, 4 µL of extracted sample was injected onto UPLC- QTOF-MS system equipped with an ESI interface (LC: Waters Acquity UPLC system; Xevo™ QT of Waters MS Technologies, Manchester, UK) operating in both negative and positive ion mode. Chromatographic separation was carried out on an ACQUITY UPLC BEH C18 column (100 mm ×2.1 mm, 1.7 μm). The column and autosampler were maintained at 40 ^◦C and 10 ^◦C, respectively. All analyses were acquired using leucine enkephalin for lock mass calibration to ensure accuracy and reproducibility, at a concentration of 0.4 ng/L, in 50/50 ACN/H2O with 0.1% v/v formic acid. The MS conditions were set as follows. Capillary voltage for negative mode: +1.5 keV; Sampling cone voltage: 35 V;

Capillary voltage for positive mode: +2.0 keV; Sampling cone voltage:

37 V; Extraction cone voltage 4 source temperature: 100 ^◦C; desolvation temperature: 450 ^◦C; cone gas flow: 30 L h⁻¹; desolvation gas flow: 850 L h⁻¹; collision energy: 6 eV, and for MS/MS spectra collision energies were set from 25 till 50 eV; detection mode: scan range was from 50–1000 m/z. In order to run the samples in the column, the mobile phase consisted of 95% water: 5% acetonitrile: 0.1% formic acid (phase A), and 0.1% formic acid in acetonitrile (phase B) were used. The solvent gradient program was conditioned as follows: 100–50% solvent A over the first 8 min, 50–0% solvent A over 1 min and return to the initial 100% A in 3.5 min, and conditioning at 100% A. The flow rate was set at 0.5 mL min⁻¹. The scans were repeated for 15 min in a single run.

Fig. 1. Schematic of the electrode design and incorporation into the electrochemical cell. (a) LIG printed on PES polymer; (b) LIG on PES with attached copper wires;

(c) electrochemical cell with LIG-PES.

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2.5. Detection and quantification of H2O2

LIG electrodes with surface areas of 4, 6, and 8 cm²were connected to a power source and immersed in an aqueous solution (70 mL, 0.05 M Na₂SO₄) in a 100 mL beaker with continuous stirring. Each test was performed in the dark. Voltages of 1.5, 2.0 and 2.5 V were applied to the cell and 2 mL samples were collected from the solution at 0, 2, 4 and 6 h, respectively. Immediately after the sample was collected, one tablet of DPD reagent [63,64] was added and the sample was shaken vigorously for 2 min. The sample was then diluted with Milli-Q water (5 mL) and the absorbance measured in the UV–visible spectrophotometer (UV-2550 Shimadzu) at 515 nm.

2.6. Characterization of LIG electrodes

Morphological characterization of the top surface of the synthesized LIG on PES film before and after use was observed using a scanning electron microscope (SEM) (Quanta 200, FEI, USA) with an accelerating voltage of 5 kV. Samples were sputter-coated with a thin layer of gold to ensure adequate sample surface conductivity. Elemental analysis of the LIG surfaces was conducted with energy-dispersive x-ray spectroscopy (EDS) analysis. The Raman measurements were done on a Horiba Lab- Ram HR evolution micro-Raman system, equipped with a Synapse Open Electrode CCD detector. The excitation source was a 532 nm laser with a power on the sample ranging from 0.6 mW to 3 mW. Typical exposure times varied from 10 to 120 s. The laser was focused with an x50 objective to a spot of about 1.3 µm. The measurements were taken using a 600 g mm⁻¹grating and a 100 µm confocal microscope hole. X-ray photoelectron spectroscopy (XPS) data were collected using an X-ray photoelectron spectrometer ESCALAB 250 ultrahigh vacuum (1 ×10⁻⁹ bar) apparatus with an AlKα X-ray source and a monochromator. The X- ray beam size was 500 μm and survey spectra was recorded with pass energy 150 eV and high energy resolution spectra were recorded with pass energy 20 eV. The spectral components of C1s, N1s, O1s and S2p signals were found by fitting a sum of single component lines to the experimental data by means of nonlinear least-squares curve-fitting. To correct for charging effects, all spectra were calibrated relative to a carbon 1 s peak positioned at 285.0 eV.

2.7. Specific energy consumption calculation

In order to estimate the cost and amount of consumed energy during the electrochemical process, the specific energy consumption (E) for the degradation of MB and CBZ was calculated with eqns. (1)-3,

Q=I×tEC (1)

E=Q×V

MEC (2)

ECost=Q×V×P (3)

where Q is the electrical charge (in Coulombs) calculated from eqn. (6), I is the current (A), tEC is the electrolysis time (h), E is the energy consumed per unit of mass of the degraded compounds, V is the voltage applied (V), MEC is the mass of the pollutant degraded in the electrolytic system, Ecost is the energy cost of the electrolytic system per unit mass of degraded compound, and P is the current price of electricity for a household in Israel (USD/kWh).

3. Results and discussion

3.1. Preparation of the LIG electrodes and the electrochemical cell An optimized PES substrate with LIG lased on to its surface was fabricated by testing various power, speed, and PPI settings of the laser.

Multiple test lasings were performed and optimal laser settings were

chosen based on the conductivity of the LIG. Optimal power was 2.5% at a scan rate of 25% and at 70 PPI and resulted in LIG with the lowest resistance (~50 Ω) with respect to other power applied, and thus the highest conductivity of ~10 S/m. The LIG was characterized using SEM, XPS, and Raman spectroscopy (Fig. 2a-c). SEM images revealed a porous, foam-like structure (Fig. 2a). Raman spectra showed characteristics of graphene and included a D peak at ~1350 cm⁻¹and a G peak at ~1580 cm⁻¹and a 2D peak at ~2700 cm⁻¹(Fig. 2b). The presence of the signature G peak, along with the 2D peak, support the existence of single layer graphene sheets while the presence of the D peak indicates its surface has defects [65]. Thus, the I2D/IG peak intensity ratio indicated multilayer graphene, with significant defects indicated by ID/IG, but typical of LIG [66]. Additionally, XPS of the surface shows an elemental composition of mainly carbon (98%) and oxygen (2%) (Fig. 2c). The characterization of the surface indicated that the LIG was successfully made as previously reported [67].

3.2. Electrochemical cell testing

For measurement of the generated reactive oxygen species, H2O2

concentrations were monitored over time with DPD reagent as shown elsewhere [52]. H2O2 served as an indicator of the oxidation potential of the advanced oxidation process since, comparatively, it is the only stable activating oxygen species readily generated among its counterparts [64]. As voltages increased from 1.5 to 2.5 Vs, the amount of generated H₂O₂increased (Fig. 3a). After 6 h, the amount of H₂O₂generated reached ~0.7 ppm at 1.5 V, ~1.2 ppm at 2.0 V and ~2 ppm at 2.5 V (Fig. 3a). Of all tested voltages, 2.5 V produced the largest amount of H2O2 suggesting a greater amount of formed reactive oxygen species present for organic contaminant degradation. Therefore, 2.5 V was used for further testing.

The area of the electrode, too, was varied using a constant electrical potential of 2.5 V. As the electrode’s active area increased from 4 cm²to 8 cm², the amount of H2O2 increased from ~1.4 to ~2.2 ppm in 6 h (Fig. 3b). Although larger electrode areas might lead to even more effective generation of reactive oxygen species, electrodes of 8 cm²(with active areas of 7 cm²) were chosen for further testing because of the convenient fit in the 100 mL beaker, where 70 mL experimental solution practically immersed the electrode without wetting the connections.

3.3. Electrochemical degradation of methylene blue

Electrochemical degradation of MB was monitored over time using 5 ppm of MB, 7 cm²LIG electrodes, and 2.5 V. A gradual decolouration of MB was observed with time and a visually decolorized solution was obtained after 6 h (Fig. 4a). The decolouration was measured quanti- tatively using UV spectrometry (Fig. 4b). The initial solution of MB showed an absorbance of ~1.1 and an absorbance of less than ~ 0.05 after 24 h indicating the complete decolorization of the dye (i.e., 95%

dye removal). The final solution pH was 4.4 and the decolored solution remained unchanged in color despite the addition of base to bring the solution back to neutral conditions.

Note, since pH can affect the color of MB [68], the color of 5 ppm MB solutions was analyzed at various pH levels. At pH 13.3, 7.7, and 1.3 the color of the 5 ppm MB solution remained despite being altered to various shades of blue (Fig. 4c). The persistence of blue coloring confirmed that the decolorizing of the MB seen during the electrochemical oxidation experiments was not due to pH variations.

To determine the effect of applied voltage on MB degradation, the electrochemical cell was operated at 1.5 and 2.0 V and the results were compared to the experiments performed at 2.5 V (Fig. 4d). At voltages less than 2.5 V, MB removal was less effective with degradation values reaching ~70% and ~76% at 1.5 V and 2.0 V, respectively, after 6 h compared to the ~82% seen at 2.5 V within the same timeframe. For all voltages, the amount of MB removed continued to increase even after the applied voltage was turned off from 6 to 24 h (i.e., MB removal

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increased to ~ 80, ~ 90 and ~ 96% for 1.5, 2.0, and 2.5 V, respectively).

Thus, we suspected that MB was being adsorbed onto the LIG electrodes.

To test this, a control experiment was performed in which the LIG electrodes were inserted into the electrochemical system with no applied voltage (Fig. 4d, Control 1). The amount of adsorption of MB on LIG-PES was ~10% after 6 h and ~18% after 24 h. This is similar to the amount of MB adsorbed onto the LIG electrodes where voltages are applied during the first 6 h of experimentation and then removed.

An additional control experiment was performed to check whether exogenously added H₂O₂would degrade methylene blue in the absence of an applied voltage (Fig. 4d, Control 2). When 2 ppm of H2O2 was added without an applied voltage, the amount of MB degradation was similar to the previous control experiment, ~10% removal after 6 h and

~15% removal after 24 h. These results show that H2O2 alone is not sufficient to degrade MB and the generation of reactive oxygen species such as OH^.radicals are necessary for degradation.

Lastly, to benchmark the performance of the LIG electrodes under optimal conditions (i.e., 2.5 V) to that of Fenton’s chemistry, reaction rate constants were calculated using Eqn. S1 for the both degradation profiles under similar conditions (i.e., ~2 ppm of H2O2; Fig. 4d vs.

Fig. S3). Upon normalizing the measured rate constants to the catalyst concentration, the LIG electrodes have a rate constant more than an order of magnitude greater than that for Fenton’s chemistry (i.e., 47.7 Lg⁻¹hr⁻¹vs. 4.0 Lg⁻¹hr⁻¹, respectively). This suggests that the electrochemical degradation of organic material via LIG electrodes is a much more efficient catalytic process than that of Fenton’s chemistry.

Again, the color removal from the MB solution was observed and monitored by spectrophotometric analysis (Fig. 4a-b). The most

plausible explanation of the color removal is the electrochemical oxidation of MB through in-situ generation of reactive oxygen species such as OH^.radicals, which, in turn, react with MB mineralizing it to CO2

and H2O as described by Panizza et al. [69]. Electrogeneration of OH^. radicals by water electrolysis can be seen in eqn. (3) below. Possible reaction pathways are as follows [70]:

The one-electron process:

H2O→OH⋅(aq) + (H⁺+e⁻)E⁰=2.38V (4) The two-electron process:

2H2O→H2O2+2(H⁺+e⁻)E⁰=1.76V (5)

Eqs. (3)-(4) indicate that the electrolysis of water can result in the generation of H2O2 or OH^.radicals, which are competing reactions taking place in solution. OH^.radicals might then react immediately with MB or H2O2 to form other reactive oxygen species such as HO2. (Eq. (5)).

Eventually, the reaction of these highly reactive radical oxygen species with MB can result in its complete mineralization to CO2 and H2O (as suggested in Eq. (6)).

H2O2+OH⋅→HO2⋅+H2O (6) C16H18ClN3S^51OH̅→16CO2+6H2O+H2SO4+3HNO3+HCl (7)

Oxidation is reported to be the cause of decolourization and, in some cases, the mineralization of MB to CO2 and H2O Eq. (6), [69]) with several reports showing the decolorization of MB under various oxidizing conditions and applied voltage potentials. Examples include nanofine sol of TiO2/UV, Fenton’s oxidation (Fe²⁺/H2O2) and heated Fig. 2. Characterization of newly prepared LIG electrodes. (a) Top view SEM image; (b) Raman spectrum of the PES-LIG surface; (c) XPS spectrum of the PES- LIG surface.

Fig. 3. Electrochemical H2O2 generation using LIG electrodes. (a) LIG electrode size 8 cm²(active area 7 cm²) tested with 1.5, 2.0, and 2.5 V for 6 h in Na2SO4(aq)

0.05 M. (b) LIG electrodes with surface area of 5, 7, 9 cm²(active area of 4, 6, and 8 cm²) at 2.5 V.

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persulphate in aqueous solution [71–73] among others which use UV [74,75] and porous PANI-CNT electrodes [76]. In light of these studies, the presently tested PES-LIG electrodes are proposed to generate

reactive oxygen species, which can oxidize organic pollutants such as MB via the reaction pathways outlined in Eqs. (3)-((6). Electrochemical oxidation can be described by either direct anodic – where the pollutant Fig. 4. Electrochemical degradation of MB. (a) MB solutions collected over time; experimental conditions included 5 ppm initial concentrated solution of MB, with LIG electrodes of 7 cm²at 2.5 V. (b) UV-Spectrometric analysis of experimental solution at different time intervals measured at the wavelength of 656 nm. (c) MB solutions at various pH. (d) Normalized concentration of MB over time for various applied voltages (i.e., 0.0 V – 2.5 V); initial concentration of MB =5 ppm, MB concentrations determined with UV–vis at 656 nm. Voltages only applied for the first 6 h of the experiment. Controls included: operation at 0.0 V (control 1) and operation at 0.0 V with 2 ppm H2O2 added exogenously (control 2). The data presented here is average of the three replicates.

Fig. 5. Characterization of LIG electrodes after use in degradation of MB. SEM images (top view) of the (a) cathode and (b) anode. Raman spectrum of the (c) cathode (d) and anode.

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is destroyed at the anode surface – or indirect oxidation – where electrochemically produced mediators (e.g., H₂O₂, HSO₄⁻, Cl₂) oxidize concomitant species in solution [77]. Considering Chiang et al. showed that both oxidation pathways can co-exist during the electro oxidation of aqueous effluents [77], the decoloration/oxidation of MB seen in this study is most likely achieved by a combination of both direct and indirect oxidation.

The total organic carbon (TOC) of the experimental solution was also measured to understand the extent of MB mineralization. TOC results show that ~40% TOC reduction was achieved after 6 h for the 5 ppm MB solution under an applied voltage of 2.5 V. This result is supported by the similar observations of Qian Zhang et al. who used high photon flux UV irradiation to degrade MB [78]. The initial oxidation reactions will affect the pi electron conjugation and electronic resonance in the aromatic rings of the MB which affects the color, but the change in TOC can only be observed when a compound is completely degraded and mineralized to non-organic carbon like CO2. Similar trends are reported in the literature when zero valent iron is used to degrade MB [79].

To identify any chemical or physical changes that might have occurred on the surface of the electrodes during experimentation, each used electrode was characterized before and after the experiment.

Oxidation reactions occurring at the anode can affect the structure and chemistry of the LIG surface [80]. SEM revealed that after the electrodes were used in the MB experiments, the LIG anode and cathode had a similar porous foam-like structure to the freshly prepared LIG (i.e., Fig. 5a,b vs. Fig. 2a). Also, like freshly prepared LIG, the Raman spectra of the used samples showed characteristics of graphene and included a D peak at ~1350 cm⁻¹and a G peak at ~1580 cm⁻¹and a 2D peak at

~2700 cm⁻¹(i.e., Fig. 5c,d vs. Fig. 2b). The 2D/G ratio of both of the electrodes are similar indicating similar quantities of single and multilayer graphene content [66]. However, the D/G ratio is slightly different for cathode and anode. The D/G ratio is higher in the anode, indicating more defects than the cathode and might indicate slight surface oxidation. Also, both electrodes were observed to have a higher D/G ratio than the unused LIG, which might indicated that the graphene component’s defects increase during usage [65].

3.4. Electrochemical degradation of carbamazepine

Electrochemical degradation of CBZ was performed using the same electrochemical apparatus as for MB. Using LIG electrodes with a surface area of 7 cm², an electrolyte solution containing 2 ppm CBZ was tested

using an applied voltage of 2.5 V for 6 h. Here too, after 6 h, the voltage was turned off and the solution left to stand for 24 h (Fig. 6). The UPLC full spectrum raw data are shown in Fig. S4. A gradual decrease in the normalized concentration of CBZ was observed using HPLC analysis to estimate the concentration, and MS to identify the compound. The normalized concentration of CBZ (m/z =237) rapidly decreased around 64% during the first 6 h when 2.5 V was applied, and then more slowly from 6 to 24 h. At 24 h, the total concentration of CBZ was found to have decreased by 85%. The same experiment was performed using an applied voltage of 1.5 and 2.0 V, and the concentration of CBZ decreased by ~59% and ~62% after 6 h, respectively. No significant decrease in concentration was detected for the control (i.e., no voltage applied) with the PES-LIG electrodes partially immersed in the solution and neither when 2 ppm of H2O2 was exogenously added to the solution. After 24 h, the concentration of CBZ decreased by ~20%, attributed to CBZ adsorption onto the electrodes and H2O2 oxidation. Thus, similar to the MB experiments described above, electrical effects are required for efficient CBZ removal including the generation of short lived OH^.radicals or other highly reactive oxygen species.

To provide some evidence that might lead to a clearer understanding of the reaction mechanism, the degradation products of CBZ were quantified and identified with UPLC-MS (Fig. 7). The main compound in the sample appeared as a sharp peak at a retention time of 5.48 min. This compound was clearly identified as CBZ (m/z =237) using MS. The area under the peaks of the chromatogram is proportional to the concentration of those analytes and comparison of t =0 h and t =6 h showed a decrease in the concentration of CBZ by ~ 64%, which is in agreement with the HPLC results. Five degradation products of CBZ were also detected by MS and the concentration of each metabolite was estimated (Fig. 7a-e). Close analysis of the UPLC-MS spectra indicated that these partially degraded CBZ products consisted of only 3.5% of the total (Table S2). Thus, the rest of the CBZ degradation products were prob- ably degraded into much smaller, undetectable CBZ compound fragments or were completely mineralized by the electrochemical degradation with the LIG electrodes.

Four of the five oxidation products of CBZ detected could be identified (Fig. 7). The first oxidation product detected with (m/z =226.13) at a retention time of 5.92 min (Table S2) and had an increasing concentration over 6 h. Based on previous CBZ degradation studies by Sun et al. this compound is most likely iminostilbene-10,11-dihydrodiol (Fig. 7b) [81]. The second oxidation product detected with (m/z = 251.121) at retention time 2.48 min was identified as carbamazepine-10,11-epoxide (Fig. 7c, [61]) and has a higher mass than the parent compound. The third oxidation product with (m/z =208.2) at retention time 5.9 min had an increasing concentration over 6 h as well.

This was designated as acridine-9-carboxaldehyde (Fig. 7d) – a degradation product of CBZ reported by Vogna et al. [16]. The fourth oxidation product detected had a retention time of 2.5 min with m/z = 180.08 and was identified as acridine (Fig. 7e) [16]. The last detected oxidation product yielded a major peak in the UPLC spectrum with (m/z

=129.12). This peak is unreported in literature and insufficient evidence prohibits the proposal of its molecular structure.

Several suggested competing reactions can occur inside the solution during the electrochemical oxidation of CBZ (eqns. (3)-5 & 7, [70,82]).

These reactions also describe how H2O2 and OH⋅ are produced in a cycle starting from water splitting (eqn. (3)). Such generated oxidizing agents more than likely react with CBZ in the presence of applied voltage to degrade a pollutant compound to its mineralized products. Since UPLC could identify some of the degradation products of CBZ, the probable mechanism of degradation and mineralization is presented in Fig. 8 below.

2HO2⋅→H2O2+O2 (8)

The degradation pathway is initiated with hydroxyl radical generation as reported elsewhere [83–86]. The CBZ starting compound (m/z = 237) is attacked by a hydroxyl radical resulting in an unstable radical Fig. 6. Electrochemical degradation of CBZ using LIG electrodes. Initial con-

centration of CBZ =2 ppm determined from HPLC. CBZ concentration monitored for 24 h with 2.5 V applied for the first 6 h (black line) and 0.0 V applied (red line, includes exogenous addition of 2 ppm H2O2). The data presented here is average of three replicates.

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compound, which in turn forms a carbamazepine epoxide (m/z = 251.12) by hydration of the C10-C11 double bond of the parent compound [83]. This epoxide can be the precursor for two products which were detected in UPLC with (m/z =226.13) and (m/z =208.2). The compound (m/z =208.2) – identified as acridine-9-carbaldehyde – can degrade into acridine (m/z =180.08) and is likely formed from the contraction of the carbamazepine epoxide ring, hydroxylation, and the

loss of a CONH2 group [83]. Via a ‘ring opening’ mechanism of acridine, smaller compound fragments can result consisting of saturated or un- saturated carbon chains having a much smaller mass than the initial compound and are challenging to detect by UPLC [84]. These smaller compound fragments or completely mineralized CBZ products account for most of the degradation products (61.5%). Although differing from contemporary mechanistic interpretations, the major pathway of CBZ Fig. 7.Proposed CBZ oxidation and degradation pathway. Detected degradation products of CBZ from UPLC-MS: (a) CBZ (m/z =237), (b) iminostilbene-10,11- dihydrodiol, (c) carbamazepine-10,11-epoxide, (d) acridine-9-carbaldehyde, (e) acridine.

Fig. 8.Characterization of LIG electrodes after use in electrochemical degradation of CBZ. SEM images (top view) of the (a) cathode (b) and anode. Raman spectrum of the (c) cathode (d) and anode.

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degradation reported by Vogna et al. [16] remains consistent despite the variations of formed intermediates between cited studies – degradation intermediates can vary with reaction conditions (i.e., applied H2O2/UV dosages) [83,86].

Here too, each electrode used for the electrochemical oxidation of CBZ was characterized before and after the experiment to determine any chemical or physical changes that might have occurred on their surfaces.

The LIG electrodes after a single electrochemical degradation experiment of CBZ revealed porous foam-like structures for both the cathode and the anode, and were similar to the surface of the unused LIG (Fig. 8a, b vs. Fig. 2a), although qualitative observations indicated that the LIG electrodes might have become more porous. Raman spectra showed characteristics of graphene and included a D peak at ~1350 cm⁻¹and a G peak at ~1580 cm⁻¹and a 2D peak at ~2700 cm⁻¹(Fig. 8c,d). The difference in the 2D/G ratio for cathode and anode was less significant indicating no significant structural or chemical changes occurred as a result of the electrochemical tests performed.

3.5. Energy consumption and cost analysis

Energy consumption and costs were calculated using Eqs. (1)-(3).

The calculated energy consumed for MB decolourization was 38.8 kWh kg⁻¹yielding a total cost of ~6.71 USD kg⁻¹given the current price of electricity in Israel (i.e., 0.173 USD kWh⁻¹as of December 2020 [87]).

Similarly, using TOC measurements, which indicate complete mineralization of MB, the energy consumed was 79.6 kWh kg⁻¹TOC corresponding to a net energy cost of ~13.77 USD kg⁻¹. Comparison to other literature examples include electrochemical oxidation of Acid violet 7 textile dye using Pt/Ir electrodes (max. 317.9 kWh kg⁻¹COD [88]) which produced an estimated energy cost of ~55 USD kg⁻¹for the entire system – approx. 8×greater than that estimated for the LIG electrode system – and the anodic oxidation of MB by BDD electrodes (245 kWh kg⁻¹[89]) with a system energy cost of ~42.39 USD kg⁻¹(6×greater than the LIG electrodes).

For CBZ degradation, the energy consumed by the LIG electrodes was 80.6 kWh kg⁻¹with an energy cost of ~13.49 USD kg⁻¹, ~10% less than a similar study that used activated carbon electrodes for pollutant degradation (89.3 kWh kg⁻¹, 15.45 USD kg⁻¹) [90]. However, the LIG electrodes were less efficient than BDD electrodes reported by Panizza et al. [91], which degraded an organic model compound with a much simpler chemical structure than CBZ (i.e., phenol) in wastewater treatment at an energy rate and cost of 32 kWh kg⁻¹and ~5.54 USD kg⁻¹, respectively – note: as compound structure complexity increases, longer treatment times and, therefore, higher energy cost might occur. How- ever, considering that BDD electrodes are about $310 [92] compared to the LIG electrodes produced in this study which are estimated to cost

<$0.01, the LIG electrodes might be considered as a disposable electrode solution, despite the higher energy costs than that of the BDD electrodes.

4. Conclusion

This work describes the development of PES-LIG electrodes for the electrochemical oxidation of micro pollutants. LIG-derived electrodes can degrade/removed ~62% of CBZ after 6 h and ~80% of CBZ after 24 h and ~82% to ~95% of MB after 6 h and 24 h, respectively, when exposed to 2.5 V. As found, the process was energy efficient for micro- pollutant degradation as compared to other methods. And, it does not require the addition of chemicals for the degradation process meaning no chemical sludge is produced during the process as opposed to Fenton oxidation processes. LIG electrodes are energy efficient and much less costly than alternatives like BDD electrodes. LIG might be useful in the future to treat micro-pollutants in water considering that this advanced oxidation process can be adapted to accommodate various environmental technologies. However, due to the oxidation of the electrodes themselves, applications that permit the facile replacement of the

electrodes should be considered. Future research based on this work could include fabricating LIG electrodes on porous membrane surfaces, where filtration and AOPs can occur simultaneously and pollutant residence time in the LIG electrode material can be considered.

Additionally, it is crucial to test LIG electrodes under real or relevant environmental conditions. For example, CBZ has typical concentrations of ~ 5 μg L⁻¹in groundwaters and MB exists in concentrations of ~50 - 75 mg L⁻¹in textile wastewater streams [93,94]. Moreover, future studies should consider the effect of concomitant contaminants such as:

(1) naturally occurring organic matter and/or (2) sulfate and chloride ions [94,95]. Naturally occurring organic matter can foul the carbon electrode surface via irreversible adsorption and can compete for reactive oxygen species with organic micropollutants. Sulfate ions can serve a beneficial purpose by increases electrochemical degradation performance while chloride ions can have a more hazardous effect via generating undesired halogenated disinfection by products [95–97].

Therefore, future studies using this low-cost technology and optimized conditions should be conducted. Moreover, future studies should also test system efficiency in complex water compositions to assess the future success of this technology for wastewater treatment.

Declaration of Competing Interest

The authors declare the following financial interests/personal re- lationships which may be considered as potential competing interests:

Ben Gurion University of the Negev currently owns intellectual property on some aspects of LIG, licensed to a company in which CJA owns stock, although he is not an officer or director.

Acknowledgments

We are grateful to the United States-Israel Binational Science Foun- dation (BSF Grant No. 2014233), and Ministry of Science and Tech- nology of the State of Israel and the German Federal Ministry of Education and Research (BMBF, PTKA), BMBF 02WIL1487 for financial support. CDP thanks the Fulbright scholar program for support. CT is grateful to the Department of Science and Technology (DST), New Delhi, India, for awarding the INSPIRE Faculty Award (DST/INSPIRE/04/

2019/000925).

Notes

Ben Gurion University of the Negev currently owns intellectual property on some aspects of LIG, licensed to a company in which CJA owns stock, although he is not an officer or director.

Supplementary materials

Supplementary material associated with this article can be found, in the online version, at doi:10.1016/j.ceja.2021.100195.

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