Electrochemically activated peroxymonosulfate for the abatement of chloramphenicol in water: performance and mechanism

Abstract

In this study, electrochemically activated peroxymonosulfate (EC/PMS) with a sacrificial iron electrode was used for the removal of chloramphenicol (CAP) from water. Compared to electrolysis alone, peroxymonosulfate (PMS) alone, and Fe²⁺/PMS, EC/PMS significantly enhanced the CAP degradation. Various parameters, such as the applied current, electrolyte concentration, and PMS dose, were investigated to optimize the process. In addition, acidic conditions facilitated the CAP degradation. The presence of Cl⁻ slightly enhanced the CAP degradation, while both HCO₃⁻ and NO₃⁻ exhibited an inhibitory effect on the CAP degradation. The floccules were also analyzed after the reaction by XPS and XRD. Quenching experiments indicated that both sulfate radicals (SO₄^●−) and hydroxyl radicals (•OH) were responsible for the CAP degradation. In addition, the degradation products were identified by LC/TOF/MS, and the degradation pathways were proposed accordingly. These results indicated that EC/PMS is a promising treatment process for the remediation of water polluted by CAP.

Introduction

Various kinds of antibiotics have been produced and applied worldwide for the treatment of human and animal diseases caused by microbial infectious diseases. As a result, a large amount of antibiotics has been discharged into the environment via effluent from hospitals and wastewater treatment plants (WWTPs), pharmaceutical industry wastewater release, improper discharge of treated landfill leachate, etc. (Kümmerer 2009). Residual antibiotics in the water environment pose potential risks to aquatic organisms as well as to human health, and they can also cause the formation of antibiotic-resistant bacteria (ARB) and antibiotic resistance genes (ARG) (Danner et al. 2019; Kovalakova et al. 2020). Due to excessive consumption and incorrect treatment methods, antibiotic pollution in the water environment has raised widespread concern and become a common environmental problem (Yan et al. 2015). As one of the commonly used antibiotics, chloramphenicol (CAP) is a broad-spectrum antibiotic for both humans and animals owing to its high activity against many gram-positive and gram-negative bacteria (Wu et al. 2017). It has been reported that a considerable proportion of CAP is directly excreted in the aquatic environment via urine or feces after its intake by humans or animals (Zhou et al. 2017). Moreover, many antibiotics, including CAP, are stable in water and are resistant to traditional wastewater treatment owing to their physical and chemical properties (Cao et al. 2020; Zhou et al. 2009). Accordingly, nano- to micrograms CAP has been frequently detected in aquatic environments. In Germany, CAP concentration was up to 0.56 μg/L and 60 ng/L, as detected in the effluent of one WWTP and in surface water, respectively (Hirsch et al. 1999). As detected in the Pearl River of China, the concentration of CAP was as high as 266 ng/L and 187 ng/L in the high and low water seasons, respectively (Xu et al. 2007). Therefore, effective techniques are urgently needed for efficient CAP removal.

Many techniques have been extensively adopted for eliminating CAP from water, such as adsorption (Fan et al. 2010), ferrate (Zhang et al. 2021), and solar photoelectro-Fenton (Garcia-Segura et al. 2014). Among these techniques, advanced oxidation processes (AOPs) are promising methods for destruction of many organic contaminants in water, which are based on the generation of highly oxidizing hydroxyl radicals (•OH). Recently, sulfate radical (SO₄^●−)–based advanced oxidation processes (SR-AOPs) have attracted much interest recently due to their excellent performance in degrading various refractory organic contaminants from water (Oh et al. 2016). In SR-AOPs, organic contaminants can be decomposed to CO₂ and H₂O as a result of SO₄^●−, which has a strong oxidation ability (E₀ = 2.5~3.1 V), high pH adaptability (2~10), and a long lifetime (30~40 μs) (Qi et al. 2017). Generally, SO₄^●− can be generated mainly via heat, UV irradiation, and transition metal activation of persulfate (PS) and peroxymonosulfate (PMS). PMS is considered to be a green oxidant that is easily activated for a variety of industrial and consumer applications involving disinfection (Tan et al. 2018). The oxidation ability of PMS at room temperature is low (E₀ = 1.82 V) (Wang and Chu 2012), but it can still be activated by various methods to produce SO₄^●−. Over the past few years, ferrous-activated PMS (Fe²⁺/PMS) has been adopted as an effective tool for SO₄^●− production for the removal of organic contaminants. Owing to its low cost, nontoxicity, and high activity, Fe²⁺ is always selected as an ideal activator of PMS or PS (Huang et al. 2021; Zou et al. 2013). Furthermore, owing to its asymmetrical structure (HO-O-SO₃⁻) and lower bond dissociation energy, PMS is considered to be more readily activated by Fe²⁺ than PS (Oh et al. 2016).

However, like the Fenton reaction, the Fe²⁺/PMS process contains the disadvantage of a narrow acidic pH range because Fe²⁺ tends to transform into iron hydroxides at neutral or alkaline pH; this conversion may hinder the transformation of Fe³⁺ to Fe²⁺ (Song et al. 2020). Furthermore, the weak reversibility of the transformation of Fe³⁺ to Fe²⁺ reduces the activation efficiency of PMS and the utilization of Fe²⁺; therefore, a higher dose of ferrous salt is needed to continuously activate PMS, which may also quench SO₄^●− and produce a large amount of iron sludge (Ali et al. 2020). An effective approach to address such drawbacks is introducing an electric field into the Fe²⁺/PMS system (Brillas et al. 2009). Electrochemically activated PMS (EC/PMS) that uses Fe as the sacrificial anode is an ideal alternative for PMS activation because it can overcome the drawback of the slow Fe³⁺/Fe²⁺ cycle that occurs in the Fe²⁺/PMS process (Govindan et al. 2014). The EC/PMS process is a combination of electrochemical and Fe²⁺/PMS processes. In the EC/PMS system, soluble Fe²⁺ is continuously released to the solution from the surface of the iron electrode through anodic oxidation. Meanwhile, the reduction of Fe³⁺ occurs at the cathode (Wang and Chu 2011). Compared with the Fe²⁺/PMS process, EC/PMS can not only control the release of Fe²⁺ by adjusting the applied current but also accelerate the Fe³⁺/Fe²⁺ cycle, thus overcoming the low activation efficiency of PMS and the iron sludge problem caused by addition of excessive Fe²⁺ in the Fe²⁺/PMS process. In addition, electrolysis can also directly activate PMS to generate SO₄^●− (Wang and Chu 2011).

Herein, an electrochemically activated PMS (EC/PMS) process that uses iron as the sacrificial anode was proposed for the abatement of CAP. Several influencing factors, including the applied current, electrolyte concentration, PMS dose, initial solution pH, and common anions, were investigated to determine their effect on the CAP degradation performance. Moreover, the floccules formed during the reaction were characterized by X-ray photoelectron spectroscopy (XPS) and X-ray diffraction (XRD). The radical species responsible for CAP degradation were investigated by quenching experiments. In addition, the major degradation products of CAP were identified by LC/TOF/MS, and the degradation pathways of CAP during the EC/PMS process were proposed.

Materials and methods

Materials

Chloramphenicol (CAP, ≥98%) was purchased from Sigma-Aldrich (St. Louis, MO, USA). Peroxymonosulfate (KHSO₅·1/2KHSO₄·1/2K₂SO₄, ≥47%) was purchased from Aladdin Chemistry Co., Ltd. (Shanghai, China). Anhydrous sodium sulfate (Na₂SO₄, ≥99%) was purchased from Sinopharm Chemical Reagent Co., Ltd. (Shanghai, China). Chromatographic reagent grade acetonitrile used as the mobile phase was purchased from J. T. Baker (USA). All other chemicals used in the study were of at least analytical grade. Milli-Q water (18.2•MΩ cm) was used to prepare all the experimental solutions.

Experimental procedure

The experiments were carried out in a 250-mL undivided single-compartment glass cell by using a plate iron anode (2 cm × 4 cm) and a plate cathode (stainless steel) of the same dimensions (Fig. 1). Constant current electrolysis conditions were maintained by a DC-regulated power supply provided by Dongguan Tongmen Electronic Technology Co., Ltd. (China). The two electrode plates were fixed parallel to each other at a distance of 2.5 cm. The reaction temperature was kept at 20 ± 1 °C by immersing the reaction cell in a water bath. In a typical run, a mixed solution of CAP and electrolyte (Na₂SO₄) was added to the reactor and then stirred by a magnetic stirrer. After adding a certain dose of PMS into the reaction solution and turning on the DC power, the reaction was initiated. The initial solution pH was adjusted by H₂SO₄ or NaOH. At predetermined time intervals, the samples were withdrawn from the electrolytic cell and filtered through 0.45-μm membranes and then quenched with prefilled methanol immediately before analysis.

Fig 1

Scheme of the experimental equipment

Analysis

The concentration of CAP was determined by Shimadzu HPLC consisting of an LC-20AB pump and an SPD-20A UV detector set at 275 nm. A mixture of acetonitrile/water (60:40, v/v) was used as the mobile phase, and the flow rate was set at 0.8 mL/min. The concentration of PMS was determined by UV spectrophotometry according to the method described by Wacławek et al. (2015). The floccules were characterized by XRD (Rigaku Ultima IV) and XPS (Thermo Scientific K-Alpha) after the reaction. The CAP degradation products were analyzed using ultra-performance liquid chromatography and quadrupole time-of-flight mass spectrometry (Agilent 1290 UPLC/6550 Q-TOF) with a Waters BEH C18 column (2.1 × 100 mm, 1.7 μm). An elution gradient with two mobile phases, MQ water (A) and acetonitrile (B), was programmed as follows (min, %A): (0, 90), (1, 90), (8, 10), (12, 10), (12.1, 90), and (13, 90). The flow rate was 0.3 mL/min. The mass spectrometer was operated in negative ionization mode using an electrospray ionization (ESI) source. The spray voltage, sheath gas temperature, and sheath gas flow were set at −3200 V, 350°C, and 12 L/min, respectively. Full scan mode was set to m/z 50–500.

Results and discussion

Comparative study of different systems

We first investigated CAP degradation in five different systems: (1) EC alone, (2) PMS alone, (3) EC/PMS process, (4) EC/PMS process using stainless steel as the anode and iron as the cathode, and (5) Fe²⁺/PMS. As shown in Fig. 2, no obvious removal was observed when CAP was treated by PMS oxidation alone because PMS is stable at ambient temperature and its redox potential is very limited (E⁰ = 1.82 V) (Wang and Chu 2012). EC alone achieved 29.7% CAP removal, implying that the coagulation mechanism plays a role in CAP removal. As expected, approximately 91.0% CAP removal was achieved within 20 min in the EC/PMS system using iron as the anode and stainless steel as the cathode. During the EC/PMS process, Fe²⁺ was first electrochemically produced from the sacrificial iron anode, which could serve as the activator of PMS to generate sulfate radicals (SO₄^●−) (Eqs. (1) and (2)), which accelerate CAP degradation. Fe²⁺ was regenerated on the cathode plate according to Eq. (3), maintaining a relatively proper level of Fe²⁺ in the system. In addition, after swapping the two electrode plates (i.e., using stainless steel as the anode and iron as the cathode), we also noticed that 14.9% removal of CAP was achieved within 20 min, which indicated that electrolysis can also directly activate PMS to form SO₄^●− and/or •OH (Eqs. (4) and (5)) (Wang and Chu 2011). We should also notice that direct oxidation of CAP on anode is also expected to occur.

$${Fe}\to {{Fe}}^{2+}+2{{e}}^{-}\left({Anode}\right)$$

(1)

$${{Fe}}^{2+}+{{HSO}}_5^{-}\to {{Fe}}^{3+}+{{SO}}_4^{\bullet -}+{{OH}}^{-}$$

(2)

$${{Fe}}^{3+}+{{e}}^{-}\to {{Fe}}^{2+}\left({Cathode}\right)$$

(3)

$${{HSO}}_5^{-}+{{e}}^{-}\to {{SO}}_4^{\bullet -}+{{OH}}^{-}$$

(4)

$${{HSO}}_5^{-}+{{e}}^{-}\to {{SO}}_4^{2-}+\cdot {OH}$$

(5)

Fig. 2

CAP degradation under different systems. Experimental conditions: [CAP]₀ = 10 μM, [PMS]₀ = 0.5 mM, [Fe²⁺]₀ = 621 μΜ, current = 25 mA, [Na₂SO₄]₀ = 25 mM, pH₀ = 7

A comparison between the EC/PMS system and the ferrous ion–activated PMS (Fe²⁺/PMS) system was also conducted. The concentration of added Fe²⁺ in the EC/PMS system was calculated through Faraday’s law (Eq. (6)) with a reaction time of 20 min.

$$C=\frac{I\times t}{F\times Z\times V}$$

(6)

where C is the molecular concentration of Fe²⁺ generated in the system (mol/L), Z is the number of electrons (2 equivalents/mol), F is Faraday’s constant (96,485.3 Coulombs/mol), I is the given current (A), t is the reaction time (s), and V is the volume of electrolyte (L).

As seen, the degradation behavior of CAP in the Fe²⁺/PMS system exhibited a fast and then a slow stage along with the reaction time. The fast stage was attributed to the relatively higher initial Fe²⁺ concentration, and the slower stage was caused by the depletion of Fe²⁺ and inadequate regeneration of Fe³⁺ to Fe²⁺, which inhibited the generation of SO₄^●− (Tan et al. 2018). We also noticed that the final removal efficiency of CAP by the EC/PMS system was higher than that of the Fe²⁺/PMS system. This phenomenon could be ascribed to the different levels of Fe²⁺ use in the two systems. In the Fe²⁺/PMS system, excess Fe²⁺ competed with CAP for the generated SO₄^●− as well as rapidly consumed PMS in the initial stage and thus hindered the CAP degradation in the following stage (Eq. (7)) (Rastogi et al. 2009). However, in the EC/PMS system, Fe²⁺ was gradually released from the iron electrode, and Fe²⁺ could also be regenerated via cathodic reduction of Fe³⁺, which maintained CAP degradation over time.

$${{Fe}}^{2+}+{{SO}}_4^{\bullet -}\to {{Fe}}^{3+}+{{SO}}_4^{2-}$$

(7)

Effect of applied current

The effect of the applied current on CAP degradation by the EC/PMS system is shown in Fig. 3. As the applied current increased from 10 to 25 mA, the removal rate of CAP increased from 74.1 to 91.0% within 20 min. As the applied current further increased to 35 mA, the removal rate of CAP slightly increased from 91.0 to 93.5%. Increasing the applied current density led to more rapid Fe²⁺ production, which improved the decomposition of PMS to generate more SO₄^●−, thus enhancing CAP degradation. In addition, a higher applied current also led to a faster generation of SO₄^●− via an electron transfer reaction. Additionally, more Fe²⁺ was also generated at a higher applied current; thus, the excess Fe²⁺ may have acted as a SO₄^●− scavenger and compete with CAP for SO₄^●− (Eq. (7)) (Rastogi et al. 2009).

Fig. 3

Effect of applied current on CAP degradation. Experimental conditions: [CAP]₀ = 10 μM, [PMS]₀ = 0.5 mM, [Na₂SO₄]₀ = 25 mM, pH₀ = 7

Effect of electrolyte concentration

The effect of electrolyte concentration (Na₂SO₄) on CAP degradation by the EC/PMS system is shown in Fig. 4. As the Na₂SO₄ concentration increased from 5 to 25 mM, the removal rate of CAP also increased from 81.5 to 91.0%. The electrical conductivity of the solution was directly influenced by the electrolyte concentration, and increasing the Na₂SO₄ concentration increased the conductivity of the reaction solution, resulting in a lower energy loss during the drop in ohmic resistance of the solution (Periyasamy and Muthuchamy 2018).

Fig. 4

Effect of electrolyte concentration on CAP degradation. Experimental conditions: [CAP]₀ = 10 μM, [PMS]₀ = 0.5 mM, current = 25 mA, pH₀ = 7

Effect of PMS dose

The effect of PMS dose on CAP degradation is shown in Fig. 5a. The removal rate increased from 56.3 to 91.0% as the PMS dose increased from 0.1 to 0.5 mM, and then the removal rate of CAP decreased to 77.1% as the PMS concentration further increased to 1 mM. The increment of CAP removal rate when PMS concentration increased from 0.1 to 0.5 mM is ascribed to the fact that PMS served as the source of SO₄^●− in the EC/PMS system; thus, more reactive species were produced to degrade CAP at higher PMS concentrations. However, if the PMS concentration was too high, side reactions occurred (Eqs. (8) and (9)), leading to the self-quenching of SO₄^●− and/or •OH by excessive PMS (Shukla et al. 2010).

Fig. 5

a Effect of PMS dose on CAP degradation; b change in [PMS]_t/[PMS]₀ during the reaction. Experimental conditions: [CAP]₀ = 10 μM, current = 25 mA, [Na₂SO₄]₀ = 25 mM, pH₀ = 7

It is of interest to observe that the initial degradation rate of CAP (computed as Δc/Δt over the first 2 min, μΜ/min) was decreased from 1.64 to 0.33 μΜ/min with the increment of PMS from 0.1 to 1 mM. This observation can also be explained by the unfavorable consumption of SO₄^●− and •OH according to Eqs. (8) and (9), leading to the production of the less reactive SO₅^●− and thus slowing the CAP degradation. However, the overall removal of CAP at the end of the process was still observed at a relatively higher PMS concentration because the PMS, which served as the SO₄^●− source, was completely consumed at a relatively lower PMS concentration during the 20-min reaction (Fig. 5b).

$${{SO}}_4^{\bullet -}+{{HSO}}_5^{-}\to {{HSO}}_4^{-}+{{SO}}_5^{\bullet -}k<{10}^5{M}^{-1}{s}^{-1}$$

(8)

$$\cdot {OH}+{{H}{SO}}_5^{-}\to {{H}}_2{O}+{{SO}}_5^{\bullet -}k=1.7\times {10}^7{M}^{-1}{s}^{-1}$$

(9)

Effect of initial solution pH

The effect of the initial solution pH on CAP degradation by the EC/PMS system is shown in Fig. 6. The results showed that CAP degradation in the EC/PMS system was highly pH-dependent. As the initial pH increased from 3 to 11, the removal rate of CAP significantly decreased from 94.0 to 22.9%. The results showed that acidic conditions favored CAP degradation, and the optimum pH was 3 in this study, which is consistent with that of Fenton-like oxidation processes (Sheng et al. 1999). Because the speciation of Fe ions is highly dependent on the solution pH, as the pH increased from 3 to 11, the main forms of iron in the solution gradually changed from Fe²⁺ or Fe³⁺ to Fe(OH)⁺, FeOH²⁺, Fe₂(OH)₂⁴⁺, and then finally transformed into Fe(OH)₃ (Eqs. (10)–(12)). This transformation resulted in the dramatic decrease in the percentage of free Fe²⁺, leading to the reduction of SO₄^●− and slowing down CAP degradation (Lin et al. 2013). In addition, the reduction of iron ions in the cathode was also inhibited under alkaline conditions, which further delayed the degradation of CAP (Lakshmanan et al. 2009).

$${{Fe}}^{2+}+{{H}}_2{O}\to {Fe}{\left({OH}\right)}^{+}+{{H}}^{+}$$

(10)

$${{Fe}}^{3+}+{{H}}_2{O}\to {{Fe}{OH}}^{2+}+{{H}}^{+}$$

(11)

$$2{{Fe}}^{3+}+2{{H}}_2{O}\to {{Fe}}_2{\left({OH}\right)}_2^{4+}+2{{H}}^{+}$$

(12)

Fig. 6

Effect of initial pH on CAP degradation. Experimental conditions: [CAP]₀ = 10 μM, [PMS]₀ = 0.5 mM, current = 25 mA, [Na₂SO₄]₀ = 25 mM

Effect of common anions

Anions are ubiquitous in natural water and may affect SO₄^●−-based oxidation process. Thus, it is necessary to explore the influence of these anions on the CAP degradation performance in the EC/PMS system. The effect of common anions, i.e., chloride (Cl⁻), bicarbonate (HCO₃⁻), and nitrate (NO₃⁻), on CAP degradation by the EC/PMS system is shown in Fig. 7. Adding Cl⁻ only slightly enhanced the degradation of CAP, and the removal rate of CAP increased from 90.9 to 98.8% as the concentration of Cl⁻ increased from 0 to 10 mM. The addition of HCO₃⁻ or NO₃⁻ exerted a negative effect on CAP degradation performance, and the removal rates decreased from 90.9 to 26.6% and 66.3%, respectively, when the concentration of HCO₃⁻ or NO₃⁻ increased from 0 to 10 mM. The inhibition by HCO₃⁻ and NO₃⁻ could be ascribed to their scavenging effect on SO₄^●− and/or •OH as the following reactions (Eqs. (13)–(16)) (Ghauch and Tuqan 2012; Keen et al. 2012). However, in the presence of Cl^-, Cl^- reacted with SO₄^●− to form Cl^● (Eq. (17)), which then reacted with Cl⁻ to generate Cl₂^●− (Eq. (18)). The reactivities of Cl^● and Cl₂^●− with CAP may be similar to the reactivity between SO₄^●− and CAP (Lian et al. 2017). Furthermore, the transformation of SO₄^●− to reactive chlorine species can reduce the possibility of SO₄^●− scavenging by other SO₄^●− ions and/or PMS; this reduction in scavenging can compensate for the lower redox potential of Cl^● and Cl₂^●−.

$${{SO}}_4^{\bullet -}+{{HCO}}_3^{-}\to {{HSO}}_4^{-}+{{CO}}_3^{\bullet -}$$

(13)

$$\cdot {OH}+{{H}{CO}}_3^{-}\to {{H}}_2{O}+{{CO}}_3^{\bullet -}$$

(14)

$${{SO}}_4^{\bullet -}+{{NO}}_3^{-}\to {{SO}}_4^{2-}+{{NO}}_3^{\bullet }$$

(15)

$$\cdot {OH}+{{NO}}_3^{-}\to {{OH}}^{-}+{{NO}}_3^{\bullet }$$

(16)

$${{Cl}}^{-}+{{SO}}_4^{\bullet -}\to {{SO}}_4^{2-}+{{Cl}}^{\bullet }$$

(17)

$${{Cl}}^{\bullet }+{{Cl}}^{-}\to {{Cl}}_2^{\bullet -}$$

(18)

Fig. 7

Effect of different concentrations of a Cl⁻, b HCO₃⁻, and c NO₃⁻ on CAP degradation. Experimental conditions: [CAP]₀ = 10 μM, [PMS]₀ = 0.5 mM, current = 25 mA, [Na₂SO₄]₀ = 25 mM, pH₀ = 7

Degradation mechanism in the EC/PMS system

Characterization of floccules formed during the reaction

The composition of the floccules was analyzed after the reaction by XPS and XRD, and the results are shown in Fig. 8. According to the XPS full spectrum (Fig. 8a), the presence of Fe 2p, O 1s, and C 1s was observed in the floccules. The charge states of the elements in the formed floccules were determined by scanning the Fe 2p and O 1s orbitals. As shown in Fig. 8b, there are two spin-orbit separation peaks of the Fe 2p_3/2 and Fe 2p_1/2 core levels at binding energies of 710.4 eV and 723.9 eV, respectively (Xia et al. 2017). The peak positions of the Fe 2p_3/2 level of Fe(III), Fe(II), and Fe(0) are at 712.6 eV, 710.9 eV, and 709.9 eV, respectively. The results indicate that Fe(III), Fe(II), and Fe(0) were present in the floccules after the reaction, and their main forms may be Fe₂O₃, Fe₃O₄, and FeO(OH) (Xu and Wang 2017). When flocculation occurs during electrolysis, the cathode reduces some of the Fe²⁺ to form Fe⁰. The fine O 1s spectrum was also measured, as shown in Fig. 8c. Two intense peaks were observed at 529.6 eV and 530.9 eV. According to previous reports, O²⁻ is found near 529.7~530.1 eV, which confirms the presence of Fe₂O₃ and Fe₃O₄ in the floccules (Piumetti et al. 2015). The peak at 530.9 eV indicates the presence of hydroxides, such as FeO(OH), Fe(OH)₂, and Fe(OH)₃ (Al-Shamsi and Thomson 2013). According to both the XRD patterns (Fig. 8d) and XPS results of the floccules, the main components of the floccules are iron oxides (Fe₂O₃ and Fe₃O₄) and iron hydroxides (FeO(OH), Fe(OH)₂, and Fe(OH)₃). Therefore, in addition to coagulation and/or complexation caused by the formed floccules, PMS activation may occur, generating SO₄^●− and resulting in the promotion of CAP degradation. (He et al. 2016; Matzek and Carter 2016).

Fig. 8

a XPS full spectrum, b Fe 2p spectrum, c O 1s spectrum, and d XRD pattern of the floccules formed during the reaction

Reactive radical species identification

To verify the roles of the different reactive radical species in the EC/PMS system, quenching experiments were conducted in the presence of tert butanol (TBA) and ethanol (EtOH). EtOH (with α-hydrogen) is a well-known radical scavenger for both SO₄^●− and •OH because it can react quickly with both SO₄^●− and •OH with rate constants of (1.6–7.7) × 10⁷ and (1.2–2.8) × 10⁹ M⁻¹ s⁻¹, respectively, whereas TBA (without α-hydrogen) mainly quenches •OH because it can react more quickly with •OH ((3.8–7.6) × 10⁸ M⁻¹ s⁻¹) than with SO₄^●− ((4.0–9.1) × 10⁵ M⁻¹ s⁻¹) (Hu et al. 2020). Figure 9 shows the CAP removal efficiency in the presence of EtOH or TBA. The addition of EtOH significantly reduced the removal rate of CAP, and the removal efficiency of CAP decreased from 91.0 to 21.0% and 16.9% in the presence of 2.5 and 5 mM EtOH, respectively, after a reaction time of 20 min, suggesting that both SO₄^●− and •OH played an important role in CAP degradation. Moreover, the degradation of CAP was less inhibited by the presence of TBA, and the removal efficiency of TBA decreased from 91.0 to 65.8% and 58.6% in the presence of 2.5 and 5 mM TBA, indicating that compared to SO₄^●−, •OH played a less important role in CAP degradation in the EC/PMS at neutral pH.

Fig. 9

Effect of radical scavengers on CAP degradation. Experimental conditions: [CAP]₀ = 10 μM, [PMS]₀ = 0.5 mM, current = 25 mA, [Na₂SO₄]₀ = 25 mM, pH₀ = 7

Degradation products and transformation pathways

According to the above results, the possible reaction mechanism for the EC/PMS system was proposed, as shown in Fig. 10. In the EC/PMS system, Fe²⁺ was slowly produced from the sacrificial iron anode through electrochemical corrosion and simultaneously released into the reaction solution (Eq. (1)), which could activate PMS in solution to generate SO₄^●− and/or •OH according to Eq. (2). The absorbed Fe³⁺ on the cathode can be reduced to Fe²⁺ via an electron transfer reaction according to Eq. (3), facilitating the Fe³⁺/Fe²⁺ cycle, which reduces the wastage of the anode and maintains Fe²⁺ and radicals at appropriate concentrations (Li et al. 2018). Additionally, the adsorbed PMS on the cathode could also produce SO₄^●− and/or •OH (Eqs. (4) and (5)). The formed floccules, such as Fe(OH)₃, may also serve as coagulants, which can remove CAP through an enmeshment mechanism.

Fig. 10

Reaction mechanism and CAP degradation pathways during the EC/PMS process. Experimental conditions: [CAP]₀ = 10 μM, [PMS]₀ = 0.5 mM, current = 25 mA, [Na₂SO₄]₀ = 25 mM, pH₀ = 7

To further clarify the degradation mechanism, LC/TOF/MS was performed to identify the degradation products of CAP in the EC/PMS system. According to the results, in addition to CAP, eight major degradation products were found, and all these intermediates are summarized in Table 1. Based on the identified degradation products and previous studies, the degradation pathways of CAP were tentatively proposed in Fig. 10.

Table 1 Major DPs of CAP in the EC/PMS system identified by LC/TOF/MS

According to previous studies (Dong et al. 2017; Nie et al. 2018), DP-1 was identified as 2,2-dichloro-N-(1-(4-nitrobenzoyl)vinyl) acetamide, which is a dehydration product of CAP. During SO₄^●−-induced oxidation, hydroxylation of the aromatic ring may occur by SO₄^●− attack through electron transfer or by •OH attack through hydroxyl addition; thus, DP-2 was identified as the hydroxylated product of CAP (Dulova et al. 2017). With further oxidation, the hydroxyl group in the side chain of partial CAP transformed into carbonyl groups to yield DP-3. Reactive radicals preferentially attacked the –C–N– bond in the side chain of partial CAP, leading to the formation of DP-4 and 2,2-dichloro-acetamide, followed by further oxidation of 2,2-dichloro-acetamide to yield DP-5 (Nie et al. 2018). Under the attack of •OH, CAP was broken down into dichloro-acetic acid and DP-6. The amino group of DP-6 was further oxidized to the nitro group of DP-7 (Garcia-Segura et al. 2014). The formation of DP-8 indicated that both dehydration and the transformation of hydroxyl groups to aldehyde groups might occur at CAP. Then, these intermediate products were further degraded to smaller molecule compounds and finally mineralized to H₂O and CO₂.

Conclusion

This study showed that an electrochemically activated peroxymonosulfate (EC/PMS) system can effectively remove CAP from water, and the removal efficiency is better than that of EC alone, PMS alone, and the Fe²⁺/PMS system. Increasing the applied current and electrolyte concentration enhanced CAP degradation, and the optimal PMS dose for the removal of CAP was 0.5 mM. The removal of CAP is highly pH-dependent, and acidic conditions facilitate CAP degradation. The presence of Cl⁻ slightly promoted CAP degradation, whereas HCO₃⁻ and NO₃⁻ exhibited inhibitory effects on CAP degradation. The main components of the floccules after the reaction included iron oxides (i.e., Fe₂O₃ and Fe₃O₄) and iron hydroxides (i.e., FeO(OH) and Fe(OH)₃), as reflected by the XPS and XRD analysis results. The results of quenching experiments indicated that both SO₄^●− and •OH were responsible for the degradation of CAP. CAP degradation products were identified by LC/TOF/MS, and the primary degradation pathway was proposed.