Abstract

Recently, the advanced oxidation processes (AOPs) based on sulfate radicals (SRs) for organics degradation have become the focus of water treatment research as the oxidation ability of SRs are higher than that of hydroxyl radicals (HRs). Since the AOP-SRs can effectively mineralize organics into carbon dioxide and water under the optimized operating conditions, they are used in the degradation of refractory organics such as dyes, pesticides, pharmaceuticals, and industrial additives. SRs can be produced by activating persulfate (PS) with ultraviolet, heat, ultrasound, microwave, transition metals, and carbon. The activation of PS in iron-based transition metals is widely studied because iron is an environmentally friendly and inexpensive material. This article reviews the mechanism and application of several iron-based materials, including ferrous iron (Fe2+), ferric iron (Fe3+), zero-valent iron (Fe0), nano-sized zero-valent iron (nFe0), materials-supported nFe0, and iron-containing compounds for PS activation to degrade refractory organics. In addition, the current challenges and perspectives of the practical application of PS activated by iron-based systems in wastewater treatment are analyzed and prospected.

INTRODUCTION

In recent decades, with the increasing manufacture and the use of synthetic chemicals, a lot of wastewater containing some refractory organics is produced. Wastewater carrying these organics is usually collected by pipelines and sent to centralized wastewater treatment plants (WWTPs) for processing. However, the removal of refractory organics in the WWTPs is not complete by the biological treatment-based secondary treatment processes. It is necessary to further treat refractory organics in a more efficient way after secondary treatment processes in order to reduce the organic pollution load of the receiving water bodies.

The advanced oxidation processes (AOPs) are currently considered to be the best processes to remove refractory organics because of their strong ability to break chemical bonds of refractory organics (Fast et al. 2017). The AOPs initially investigated were based on hydroxyl radicals (HRs, HO) produced in situ to degrade organic compounds. Two oxidants, hydrogen peroxide (H2O2) and ozone (O3) are activated by ultraviolet (UV) or ferrous ions (Fe2+) to produce HRs (Deng & Zhao 2015).The main processes consisting of H2O2 or O3 are H2O2/UV, O3/UV, O3/H2O2/UV, and Fe2+/H2O2 which often called the Fenton process (Csay et al. 2014; Ghatak 2014). Based on Fenton process, photo-Fenton (Klamerth et al. 2013; Pouran et al. 2015), electro-Fenton (Nidheesh & Gandhimathi 2012; Yahya et al. 2014), sono-Fenton (Ioan et al. 2007; Ghauch et al. 2011b), and sono-photo-Fenton (Babuponnusami & Muthukumar 2011) have also been extensively studied. However, the Fenton-based processes have several defects: Both H2O2 and O3 are unstable, difficult to store and transport, and used in large quantities; the sludge generated in the reaction is difficult to treat; the optimal pH value is around 3.0, which will increase pH regulating agents (Neyens & Baeyens 2003).

In recent years, the AOPs based on sulfate radicals (SRs, SO4•−) are widely studied in degradation of organic pollutants because they overcome some shortcomings of AOPs-HRs. For instance, SRs have higher oxidation potential (E0 = 2.5–3.1 V) and longer half-life (t1/2 = 30–40 μs) than that of HRs (E0 = 1.7–2.8 V, t1/2 ≤ 1 μs), and can selectively react with unsaturated or aromatic compounds under a wide pH range of 2.0–8.0 (Olmez-Hanci & Arslan-Alaton 2013; Ike et al. 2018). PS (E0 = 2.01 V) can be activated to produce SRs in the presence of energy (UV, ultrasonic, and heat) (Guo et al. 2014; Darsinou et al. 2015; Ji et al. 2015), transition metals (Fe, Co, Ni, Cu, Ag, and Mn) (Liu et al. 2014; Rodriguez et al. 2014; Zhang et al. 2015a, 2015b; Kang et al. 2018), and carbon (Cheng et al. 2017). Among these activators, iron has been extensively studied in activating PS to generate SRs for refractory organics degradation due to its environmental friendliness, low price and excellent activation (Matzek & Carter 2016). Prior to PS, iron was even used as stand alone technology for the removal of refractory organics among them pesticides and pharmaceuticals(Ghauch 2008; Ghauch & Tuqan 2008, 2009; Ghauch et al. 2009, 2010a, 2010b,, 2011a). However, high concentrations of iron and iron-based materials were needed creating additional sludge formation. Therefore, researchers proposed iron use at trace level in combination with chemicals able to produce fugitive reactive oxygen species (ROS) (Ghauch 2015). Advantages of PS include chemical stability, ease of storage and transport, high solubility in water, and low cost, which broaden the application of PS in wastewater treatment (Tsitonaki et al. 2010). In addition, Sodium PS (Na2S2O8) and potassium PS (K2S2O8) are commonly used in research and pilot projects because they produce no secondary pollutants to water. Although high concentrations of sulfate species could be considered at risk which is not the case for PS concentration in the range of 0.1–20 mM (Liu & Wu 2018).

From the published literature available, some scholars have given an overview of the activated methods of sulfate radicals and their application in wastewater treatment (Zhang et al. 2015a, 2015b; Matzek & Carter 2016; Oh et al. 2016; Brienza & Katsoyiannis 2017; Wacławek et al. 2017; Guerra-Rodríguez et al. 2018; Xiao et al. 2018, 2019; Oh & Lim 2019). However, a detailed review about activated persulfate by iron-based materials used for refractory organics degradation has not been reported yet. Therefore, a specific and deep review on the PS activated by iron-based materials can provide state-of-the-art information for the future research of the AOPs-SRs. This paper reviews the recent advances of mechanism and application of some iron-based materials in activating PS for refractory organics degradation, and the insights towards future application of these activating methods of PS for efficient degradation of refractory organics is summarized and discussed as well.

CHARACTERIZATION METHODS OF IRON-BASED MATERIALS

In order to better understand the physicochemical properties of iron-based materials and analyze the morphology and structure of the materials, some methods of characterizing materials are necessary. There are several analytical methods such as SEM (scanning electron microscopy), TEM (transmission electron microscopy), EDS (energy dispersive system), XRD (X-ray diffraction), XPS (X-ray photoelectron spectroscopy), FTIR (Fourier transform infrared spectrometry), TGA (thermo gravimetric analysis) and BET (Brunauer-Emmett-Teller) method. The detection objects of these analysis methods are shown in Table 1.

Table 1

Functions of the instruments

SEM/TEMEDSXRDXPSFTIRTGABET
Morphology and structure Element and distribution Crystalline structure Valence state of elements Function groups The mass loss and thermal stability Surface area and pore size 
SEM/TEMEDSXRDXPSFTIRTGABET
Morphology and structure Element and distribution Crystalline structure Valence state of elements Function groups The mass loss and thermal stability Surface area and pore size 

FERROUS ION (Fe2+) AS ACTIVATOR

Mechanism of homogeneous activation

Ferrous ion is widely used as a homogeneous activator to react with PS to generate SRs. Usually ferrous sulfate (FeSO4) and ferrous chloride (FeCl2) are employed as activator to produce Fe2+ in solution. Naim & Ghauch (2016) demonstrated that chloride salt was more efficient toward PS activation than sulfate salt because of common ion effect on hand and because of the generation of chlorine radicals on the second hand responsible for additional ranitidine degradation. Table 2 lists some parameters and removal rates of refractory organics treated by Fe2+/PS system based on recent publications. The reaction mechanism of ferrous ion with PS to generate SRs is presented in Equation (1) (Jiang et al. 2013). The O–O bond of S2O82− is cleaved to form main product SO4•− and byproduct SO42− by accepting the electrons provided by Fe2+, while Fe2+ is oxidized by the S2O82− to become Fe3+. As described in Equation (2), Fe2+ is produced by the reaction between S2O82− and Fe3+ in the presence of S2O82−. Therefore, in a homogeneous system, there is a cycle between Fe2+ and Fe3+. As an oxidant, SO4•− can also be consumed by itself or excess Fe2+ and S2O82− (Equations (3)–(5)) (Kusic et al. 2011; Zhang et al. 2014), which can render a lack of SO4•− for organics degradation. On top of that, SO4•− can react slowly with H2O to form HO (Equation (6)) (Olmez-Hanci et al. 2013).
formula
(1)
formula
(2)
formula
(3)
formula
(4)
formula
(5)
formula
(6)
Table 2

Reaction conditions and treatment effects in the Fe2+/PS system

Pollutant (Abbreviation)Parameters and degradation rateRemarksReference
γ-hexachlorocyclohexane (γ-HCH) [γ-HCH] = 0.0172 mM, [PS] = 42 mM, [Fe2+] = 0.0108 mM, pH = 5.7, t = 4 h, 100% The γ-HCH oxidation rate constant (k) was relatively constant in the pH range of 3.5–9.0; however, it increased at both pHs 1.5 and 11.0. Cao et al. (2008)  
Polyvinyl alcohol (PVA) [Fe2+] :[PS] = 1:1, t = 120 min, T = 20 °C, 70%  Oh et al. (2009)  
Orange G (OG) [OG] = 0.1 mM, [Fe2+] = 1.0 mM, [Fe2+] : [PS] = 1 : 4, pH = 3.5, t = 30 min, 81.0% The following ions act as inhibitors are in sequence of NO3 <Cl <H2PO4 < HCO3Xu & Li (2010)  
Propachlor [Propachlor] = 10 mg/ L, [Fe2+] = 2.5 mM, [Fe2+] : [PS] = 1:2, pH = 4.0, T = 30 °C, t = 30 h, 60.0% Neutral or even alkaline levels leaded to strong hydrolysis of Fe2+, which had a negative effect on persulfate activation. Liu et al. (2012)  
Acetaminophen (APAP) 180 min, 18% Sodium citrate at an appropriate concentration could improve the APAP degradation. Deng et al. (2014)  
Orange G (OG) [OG] = 1.25 mM, [Fe2+] : [PS] : [OG] = 10 : 20 : 5, pH = 3.0, T = 25 °C, t = 150 min, 35.0%  Han et al. (2014)  
Toluene [Toluene] = 1 mM, [Fe2+] = 5 mM, [Fe2+] : [PS] = 1:4, pH = 7.0, t = 60 min, 70.0% Excessive dosages of PS and Fe2+ inhibited toluene degradation efficiency; the optimal pH was 4.0. Long et al. (2014)  
Carbamazepine (CBZ) [CBZ] = 0.025 mM, [Fe2+] : [PS] : [CBZ] = 10:40:1, pH = 3.0, t = 40 min, 78.0% The optimal pH was 4.22; NO3, SO42− and H2PO4 inhibited CBZ degradation rate, and Cl facilitated the decomposition of CBZ. Rao et al. (2014)  
1,4-dioxane (1,4-D) [1,4-D] = 100 mg/L, [Fe2+] = 1,000 mg/L, [Fe2+] : [PS] = 1 : 5, pH = 2.81, T = 40 °C, t = 51 h, 100% The elevated temperature (50 °C and 60 °C) caused a slight decrease in the degradation rate of 1, 4-D with the addition of Fe2+Zhao et al. (2014)  
Levofloxacin (LFX) [LFX] = 0.075 mM, [Fe2+] : [PS] : [LFX] = 1 : 10 : 1, pH = 3.0, t = 1 min, 69.0%  Epold et al. (2015)  
Chloramphenicol (CAP) The effluent of a beef cattle operation [CAP] = 0.05 mM,[PS] = 1 mM, pH = 5.4,[Fe2+] = 0.1 mM,20 injections, t = 45 min, 90.1% HCO3, NO3, NO2, H2PO4, HPO42− and HA demonstrates an adverse effect on CAP oxidation. Adding a low Cl concentration into solution can facilitate CAP degradation, but adding a high Cl concentration can inhibit CAP degradation. Nie et al. (2015)  
Reactive Black (RB5) [RB5] = 0.01 mM, [Fe2+] = 8 mM, [Fe2+] : [PS] = 1:50, pH = 7.78, t = 4 h, 98.0% The higher initial concentration of RB5, the lower decolorization efficiency; the degradation rate of RB5 decreased with increasing pH; Cl could inhibit the degradation rate of RB5. Satapanajaru et al. (2015)  
Atrazine (ATZ) [ATZ] = 20 μM, [Fe2+] = 0.4 mM, [Fe2+] : [PS] = 1:1, t = 10 min, 57.0% Low initial ATZ concentration was favorable for ATZ degradation rate. Bu et al. (2016)  
Ranitidine (RAN) [PS] = 100 μM,[Fe2+] = 100 μM, [RAN] = 28.5 μM, pH = 6.45, T = 25 °C, t = 60 min, 96% Under sequential spiking of Fe2+, higher degradation efficiency 96% of RAN is obtained. Naim & Ghauch (2016)  
Atrazine (ATZ) [ATZ] = 10 μM, [Fe2+] = 0.03 mM, [PS] = 1.0 mM, pH = 4.0, t = 15 min, 29.4%  Bu et al. (2017)  
Amicarbazone (AMZ) [AMZ] = 0.0414 mM, [Fe2+] = 0.134 mM, [PS] = 2.5 mM, pH = 3.0, t = 60 min, 35.0%  Graça et al. (2017)  
Sulfamethoxazole (SMX) [SMX] = 0.05 mM, [PS] = 4 M, [Fe2+]:[PS] = 1:1, T = 25 °C, t = 120 min, pH = 3.0, 100% The actual composition of wastewater inhibited the degradation rate of SMX. Wang & Wang (2017)  
Ketoprofen (KTP) [PS] = 0.5 mM,[Fe2+] = 0.1 mM, [KTP] = 7.87 μM, pH = 6.15, T = 25 °C, t = 60 min, 98%  The optimal [Fe2+]: [PS] ratio is 5:1. Amasha et al. (2018)  
Crystal violet (CV) [CV] = 40 mg/L, [Fe2+] = 0.3 mM, [Fe2+]: [PS] = 3:4, pH = 7.0, t = 90 min, 41.0% The optimal initial pH for degradation CV was 7.0. Chen et al. (2018)  
Bisphenol A (BPA) [BPA] = 0.0877 mM, [PS] = 4.385 mM, [Fe2+] = 0.0877 mM–0.877 mM, pH = 7.0, T = 25 °C, t = 60 min, 83.55% − 99.70% A small amount of sodium citrate in the Fe2+/PS system can lead to significant promotion of BPA removal in the rapid phase, and restore Fe3+ to Fe2+ to provide sustained activator for the slow phase. However, excessive sodium citrate can also inhibit the degradation. Gao et al. (2018)  
Chloramphenicol (CAP) [CAP] = 0.05 mM, [Fe2+] = 0.2 mM, [Fe2+] : [PS] = 2:3, pH = 3.0, t = 100 min, 27.2% Lower concentration of Cl (0.6–6 mM) slightly increased the CAP degradation rate, while higher Cl (12–36 mM) concentration inhibited CAP removal. Nie et al. (2018)  
Chlortetracycline (CTC) [PS] = 500 mM, [Fe2+] = 1,000 mM, [CTC] = 1 mM,t = 2 h, 76% All Fe2+ ions in the case of homogeneous reaction activated above 90% PS to produce SO4•− which seems to be not efficient in degrading CTC due to scavenging of sulfate radicals by itself and also by excess Fe2+ ions. Pulicharla et al. (2018)  
Trimethoprim (TMP) [TMP] = 0.05 mM, [PS] = 4 mM, [Fe2+]:[PS] = 1:1, pH = 3.0, T = 25 °C, t = 250 min, 73.4% The increasing persulfate concentration accelerated the TMP degradation rate; Under the same reaction conditions, the degradation rate of TMP in actual wastewater is 43.6%. Wang & Wang (2018)  
Acetaminophen (APAP) [APAP] = 100 mg/L,pH = 6.5, [PS] = 0.8 g/L, [Fe2+] = 0.7 mM, 90 min, 79%  Zhang et al. (2018c)  
2,4,6-tricholoroanisole (TCA) [PS] = 0.24 mM, [Fe2+] = 0.48 mM, [TCA] = 0.47 μM, pH = 3, T = 20 °C, t = 60 min, 88% The degradation of TCA by Fe3+/PS was effective at both acidic and neutral pH comparing with alkaline pH. Zhang et al. (2018b)  
Diatrizoate (DTZ) [DTZ] = 5 mg/L, [Fe2+] = 1.0 mM, [Fe2+] : [PS] = 1:10, pH = 3.0, T = 25 °C, t = 120 min, 69.0% The degradation rate of DTZ increased only 3% when pH increased from 3.0 to 9.0, but DTZ degradation rate decreased significantly at pH 11.0; more than 99% of DTZ removal is acquired at 40 °C. Shang et al. (2019)  
Triphenyl phosphate (TPHP) [TPHP] = 10 μM, [Fe2+] = 1,000 μM, [Fe2+]:[PS] = 2:1, pH = 4.0, T = 25 °C, t = 60 min, 80.0% When pH increased from 4.0 to 9.0, removal rate TPHP was almost no changed. Song et al. (2019)  
Acetaminophen (ACT) [ACT] = 0.05 mM, [Fe2+] = 1 mM, [Fe2+]:[PS] = 5:4, pH = 3.0, T = 20 °C, t = 40 min, 70.0% Initial pH (3.0–9.0) had no insignificant impact on ACT degradation; ACT degradation rate decreased with initial ACT concentration increasing from 0.025 mM to 1 mM. Wang et al. (2019c)  
Pollutant (Abbreviation)Parameters and degradation rateRemarksReference
γ-hexachlorocyclohexane (γ-HCH) [γ-HCH] = 0.0172 mM, [PS] = 42 mM, [Fe2+] = 0.0108 mM, pH = 5.7, t = 4 h, 100% The γ-HCH oxidation rate constant (k) was relatively constant in the pH range of 3.5–9.0; however, it increased at both pHs 1.5 and 11.0. Cao et al. (2008)  
Polyvinyl alcohol (PVA) [Fe2+] :[PS] = 1:1, t = 120 min, T = 20 °C, 70%  Oh et al. (2009)  
Orange G (OG) [OG] = 0.1 mM, [Fe2+] = 1.0 mM, [Fe2+] : [PS] = 1 : 4, pH = 3.5, t = 30 min, 81.0% The following ions act as inhibitors are in sequence of NO3 <Cl <H2PO4 < HCO3Xu & Li (2010)  
Propachlor [Propachlor] = 10 mg/ L, [Fe2+] = 2.5 mM, [Fe2+] : [PS] = 1:2, pH = 4.0, T = 30 °C, t = 30 h, 60.0% Neutral or even alkaline levels leaded to strong hydrolysis of Fe2+, which had a negative effect on persulfate activation. Liu et al. (2012)  
Acetaminophen (APAP) 180 min, 18% Sodium citrate at an appropriate concentration could improve the APAP degradation. Deng et al. (2014)  
Orange G (OG) [OG] = 1.25 mM, [Fe2+] : [PS] : [OG] = 10 : 20 : 5, pH = 3.0, T = 25 °C, t = 150 min, 35.0%  Han et al. (2014)  
Toluene [Toluene] = 1 mM, [Fe2+] = 5 mM, [Fe2+] : [PS] = 1:4, pH = 7.0, t = 60 min, 70.0% Excessive dosages of PS and Fe2+ inhibited toluene degradation efficiency; the optimal pH was 4.0. Long et al. (2014)  
Carbamazepine (CBZ) [CBZ] = 0.025 mM, [Fe2+] : [PS] : [CBZ] = 10:40:1, pH = 3.0, t = 40 min, 78.0% The optimal pH was 4.22; NO3, SO42− and H2PO4 inhibited CBZ degradation rate, and Cl facilitated the decomposition of CBZ. Rao et al. (2014)  
1,4-dioxane (1,4-D) [1,4-D] = 100 mg/L, [Fe2+] = 1,000 mg/L, [Fe2+] : [PS] = 1 : 5, pH = 2.81, T = 40 °C, t = 51 h, 100% The elevated temperature (50 °C and 60 °C) caused a slight decrease in the degradation rate of 1, 4-D with the addition of Fe2+Zhao et al. (2014)  
Levofloxacin (LFX) [LFX] = 0.075 mM, [Fe2+] : [PS] : [LFX] = 1 : 10 : 1, pH = 3.0, t = 1 min, 69.0%  Epold et al. (2015)  
Chloramphenicol (CAP) The effluent of a beef cattle operation [CAP] = 0.05 mM,[PS] = 1 mM, pH = 5.4,[Fe2+] = 0.1 mM,20 injections, t = 45 min, 90.1% HCO3, NO3, NO2, H2PO4, HPO42− and HA demonstrates an adverse effect on CAP oxidation. Adding a low Cl concentration into solution can facilitate CAP degradation, but adding a high Cl concentration can inhibit CAP degradation. Nie et al. (2015)  
Reactive Black (RB5) [RB5] = 0.01 mM, [Fe2+] = 8 mM, [Fe2+] : [PS] = 1:50, pH = 7.78, t = 4 h, 98.0% The higher initial concentration of RB5, the lower decolorization efficiency; the degradation rate of RB5 decreased with increasing pH; Cl could inhibit the degradation rate of RB5. Satapanajaru et al. (2015)  
Atrazine (ATZ) [ATZ] = 20 μM, [Fe2+] = 0.4 mM, [Fe2+] : [PS] = 1:1, t = 10 min, 57.0% Low initial ATZ concentration was favorable for ATZ degradation rate. Bu et al. (2016)  
Ranitidine (RAN) [PS] = 100 μM,[Fe2+] = 100 μM, [RAN] = 28.5 μM, pH = 6.45, T = 25 °C, t = 60 min, 96% Under sequential spiking of Fe2+, higher degradation efficiency 96% of RAN is obtained. Naim & Ghauch (2016)  
Atrazine (ATZ) [ATZ] = 10 μM, [Fe2+] = 0.03 mM, [PS] = 1.0 mM, pH = 4.0, t = 15 min, 29.4%  Bu et al. (2017)  
Amicarbazone (AMZ) [AMZ] = 0.0414 mM, [Fe2+] = 0.134 mM, [PS] = 2.5 mM, pH = 3.0, t = 60 min, 35.0%  Graça et al. (2017)  
Sulfamethoxazole (SMX) [SMX] = 0.05 mM, [PS] = 4 M, [Fe2+]:[PS] = 1:1, T = 25 °C, t = 120 min, pH = 3.0, 100% The actual composition of wastewater inhibited the degradation rate of SMX. Wang & Wang (2017)  
Ketoprofen (KTP) [PS] = 0.5 mM,[Fe2+] = 0.1 mM, [KTP] = 7.87 μM, pH = 6.15, T = 25 °C, t = 60 min, 98%  The optimal [Fe2+]: [PS] ratio is 5:1. Amasha et al. (2018)  
Crystal violet (CV) [CV] = 40 mg/L, [Fe2+] = 0.3 mM, [Fe2+]: [PS] = 3:4, pH = 7.0, t = 90 min, 41.0% The optimal initial pH for degradation CV was 7.0. Chen et al. (2018)  
Bisphenol A (BPA) [BPA] = 0.0877 mM, [PS] = 4.385 mM, [Fe2+] = 0.0877 mM–0.877 mM, pH = 7.0, T = 25 °C, t = 60 min, 83.55% − 99.70% A small amount of sodium citrate in the Fe2+/PS system can lead to significant promotion of BPA removal in the rapid phase, and restore Fe3+ to Fe2+ to provide sustained activator for the slow phase. However, excessive sodium citrate can also inhibit the degradation. Gao et al. (2018)  
Chloramphenicol (CAP) [CAP] = 0.05 mM, [Fe2+] = 0.2 mM, [Fe2+] : [PS] = 2:3, pH = 3.0, t = 100 min, 27.2% Lower concentration of Cl (0.6–6 mM) slightly increased the CAP degradation rate, while higher Cl (12–36 mM) concentration inhibited CAP removal. Nie et al. (2018)  
Chlortetracycline (CTC) [PS] = 500 mM, [Fe2+] = 1,000 mM, [CTC] = 1 mM,t = 2 h, 76% All Fe2+ ions in the case of homogeneous reaction activated above 90% PS to produce SO4•− which seems to be not efficient in degrading CTC due to scavenging of sulfate radicals by itself and also by excess Fe2+ ions. Pulicharla et al. (2018)  
Trimethoprim (TMP) [TMP] = 0.05 mM, [PS] = 4 mM, [Fe2+]:[PS] = 1:1, pH = 3.0, T = 25 °C, t = 250 min, 73.4% The increasing persulfate concentration accelerated the TMP degradation rate; Under the same reaction conditions, the degradation rate of TMP in actual wastewater is 43.6%. Wang & Wang (2018)  
Acetaminophen (APAP) [APAP] = 100 mg/L,pH = 6.5, [PS] = 0.8 g/L, [Fe2+] = 0.7 mM, 90 min, 79%  Zhang et al. (2018c)  
2,4,6-tricholoroanisole (TCA) [PS] = 0.24 mM, [Fe2+] = 0.48 mM, [TCA] = 0.47 μM, pH = 3, T = 20 °C, t = 60 min, 88% The degradation of TCA by Fe3+/PS was effective at both acidic and neutral pH comparing with alkaline pH. Zhang et al. (2018b)  
Diatrizoate (DTZ) [DTZ] = 5 mg/L, [Fe2+] = 1.0 mM, [Fe2+] : [PS] = 1:10, pH = 3.0, T = 25 °C, t = 120 min, 69.0% The degradation rate of DTZ increased only 3% when pH increased from 3.0 to 9.0, but DTZ degradation rate decreased significantly at pH 11.0; more than 99% of DTZ removal is acquired at 40 °C. Shang et al. (2019)  
Triphenyl phosphate (TPHP) [TPHP] = 10 μM, [Fe2+] = 1,000 μM, [Fe2+]:[PS] = 2:1, pH = 4.0, T = 25 °C, t = 60 min, 80.0% When pH increased from 4.0 to 9.0, removal rate TPHP was almost no changed. Song et al. (2019)  
Acetaminophen (ACT) [ACT] = 0.05 mM, [Fe2+] = 1 mM, [Fe2+]:[PS] = 5:4, pH = 3.0, T = 20 °C, t = 40 min, 70.0% Initial pH (3.0–9.0) had no insignificant impact on ACT degradation; ACT degradation rate decreased with initial ACT concentration increasing from 0.025 mM to 1 mM. Wang et al. (2019c)  

PS concentration and Fe2+ concentration

Some articles (Liu et al. 2012; Long et al. 2014; Rao et al. 2014; Bu et al. 2016, 2017; Nie et al. 2018; Song et al. 2019; Wang et al. 2019c) investigated whether PS alone or Fe2+ alone can remove organics at room temperature. Some results (Bu et al. 2016, 2017; Nie et al. 2018; Wang et al. 2019c) showed that neither PS alone nor Fe2+ alone can remove organics under normal conditions. However, other results (Liu et al. 2012; Long et al. 2014; Rao et al. 2014; Song et al. 2019) presented that PS can remove 8–15% of organics. One of the possible causes is direct electron transfer oxidation from PS anion to organics, and another possible cause is the photocatalysis of light or temperature towards PS (Liu et al. 2012). Long et al. (2014) found that around 10% of organics were degraded due to the presence of Fe2+ alone. The principal reason has been proposed, a series of reactions (Equations (7)–(9)) under aerobic conditions eventually lead to the generation of HO that has the ability to degrade organics (Bu et al. 2016). This shows the high concentration of dissolved oxygen in the water can promote the degradation of organics by Fe2+ alone. Caré et al. (2013) demonstrated that Fe2+ is a reductant and in some cases can remove organics however if generated in-situ from iron powder, Fe2+ is more reactive since it is accompanied with the formation of additional iron corrosion products (ICPs) responsible for the co-precipitation and sequestration of organics along with reduction.
formula
(7)
formula
(8)
formula
(9)

PS and Fe2+ concentration in solution are critical parameters to generate SO4•−, and the molar ratio of Fe2+ to PS also determines the SO4•− yield. Usually, the reaction to remove refractory organics is in two stages: at initial stage of reaction, the removal rate of organics increases with the increasing Fe2+ concentration or PS concentration; After initial stage, even the concentration of Fe2+ or PS continues to increase, the removal rate decreases. This shows that an appropriate molar ratio of Fe2+ to PS, rather than higher concentration of Fe2+ or PS, can obtain a better removal rate of organics. The most used molar ratio of Fe2+ to PS in these studies was 1:1, and followed by 1:2 according some literature (Oh et al. 2009; Liu et al. 2012; Bu et al. 2016; Wang & Wang 2017). From the reaction of Equation (1), we can see that, 1 mole of Fe2+ reacts with 1 mole of PS can produce 1 mole of SO4•−. So, theoretically, 1:1 is the optimal ratio of Fe2+ to PS. However, in fact, the optimal ratio can be greater than or less than 1:1, due to the presence of some oxidized or reduced substances in the complex sewage that can consume Fe2+ or PS. For example, SO4•− can react with excess Fe2+ or PS (Equations (4) and (5)) without considering the influence of other ions. In addition, the ratio of oxidant (PS) or catalyst (Fe2+) concentration to organics concentration is also an important factor for maintaining a better organics removal rate. Some studies (Satapanajaru et al. 2015; Wu et al. 2015; Bu et al. 2016; Nie et al. 2018) stated that the concentration of oxidant or catalyst required is proportional to the concentration of contaminants due to more pollutants require more SO4•− generated from PS activated by Fe2+.

Solution pH

The initial pH of aqueous solution is an important factor influencing the removal efficiency of refractory organics in Fe2+/PS system. Some articles (Cao et al. 2008; Xu & Li 2010; Liu et al. 2012; Niu et al. 2012; Long et al. 2014; Rao et al. 2014; Epold et al. 2015; Han et al. 2015a; Satapanajaru et al. 2015; Wu et al. 2016; Nie et al. 2018; Yu et al. 2018; Shang et al. 2019; Song et al. 2019; Wang et al. 2019c) studied the effect of pH ranging from 2.0 to 11.0 on the removal of refractory organics in Fe2+/PS system. Based on the results of these articles (Xu & Li 2010; Liu et al. 2012; Niu et al. 2012; Long et al. 2014; Rao et al. 2014; Han et al. 2015a; Satapanajaru et al. 2015; Wu et al. 2016; Nie et al. 2018), the optimum pH range was 3.0–4.0. Within such a pH range, due to acid catalysis on PS, more SO4•− could be generated from decomposition of PS, which therefore enhanced the removal rate of organics (Nie et al. 2018). When the pH was less than 3.0, refractory organics removal rate decreased because pH < 3.0 promotes the formation of the ferrous complex (Fe(H2O))2+, which reduces the amount of Fe2+ to activate PS, and thus less SO4•− can be formed. When the initial pH increased from 4.0 to 9.0, the removal efficiency of refractory organics gradually decreased. Four mechanisms are summarized to explain the reason why refractory organics degradation declined with increasing pH:

  • (1)
    When the pH is higher than 4.0, free Fe2+ turns into Fe2+ complexes (Equation (10)), which inhibits the further reaction between Fe2+ and PS to form SO4•− (Xu & Li 2010; Rao et al. 2014; Nie et al. 2018).
    formula
    (10)
  • (2)
    With pH increasing above 4.0, ferric hydroxide complexes, such as FeOH2+, Fe2(OH)24+, Fe(OH)2+, Fe(OH)3, and Fe(OH)4 formed in the solution have low activation capacity to generate SO4•− from PS, thus the efficiency of the refractory organics degradation is reduced (Nie et al. 2018). The equations of generating ferric hydroxide complexes are shown in Equations (11)–(13) (Xu & Li 2010). It can be seen from these equations that hydrogen ion is generated during the reaction and causes the decrease of pH in the solution.
    formula
    (11)
    formula
    (12)
    formula
    (13)
  • (3)
    If the pH is higher than 8.5, the present PS in the solution may react with OH to form OH (Equation (14)) in alkaline solution (Satapanajaru et al. 2015). This results in the rapid decay of SO4•− (Liu et al. 2012; Satapanajaru et al. 2015). But hydroxyl radical is also strong in its reaction with the organic compounds.
    formula
    (14)
  • (4)
    When the pH is greater than or equal to 8.4, SO4•− was found to react with HCO3 or CO32− (Equations (15) and (16)) to form HCO3 or CO3, any of which has poor oxidizing ability toward organics compared to SO4•− (Satapanajaru et al. 2015).
    formula
    (15)
    formula
    (16)

Several articles (Epold et al. 2015; Shang et al. 2019; Song et al. 2019; Wang et al. 2019c) claimed that an initial solution pH (3.0–7.0) has no significant effect on removing refractory organics. Wang et al (2019c) investigated the degradation of acetaminophen (ACT) in the Fe2+/PS system at different initial pH (3.0, 5.0, 7.0 and 9.0), and found that the degradation of ACT was around 70% after a 30-min reaction at each of initial solution pH. In the article by Wang et al. (2019c), regardless of how the initial pH changed (from 3.0 to 9.0), the pH of the solution after 1-min reaction dropped rapidly to 3.0, which demonstrates the degradation of ACT was not affected by the initial pH in Fe2+/PS system. Song et al. (2019) studied the effect of increasing pH from 4.0 to 9.0 on the degradation efficiency of triphenyl phosphate (TPhP) in Fe2+/PS process and found that degradation efficiency of TPhP was not impacted by pH changes. Epold et al. (2015) investigated levofloxacin (LFX) degradation at different pH (3.0, 5.0, 7.0 and 9.0) in Fe2+/PS system, and the degradation rates of LFX acquired were almost as same as pH from 3.0 to 7.0, while the degradation rate of LFX was decreased by about 10% at pH of 9.0. Shang et al. (2019) studied the influence of initial pH (3.0, 5.0, 7.0, 9.0 and 11.0) on diatrizoate (DTZ) degradation. It was found that DTZ degradation rate was increased to only 3% when the pH was increased from 3.0 to 9.0. However, DTZ degradation efficiency dropped significantly as the initial pH increased from 9.0 to 11.0. Shang et al. (2019) proposed that the decreased DTZ degradation at initial pH of 11.0 is due to the reduction of Fe2+ that tended to be hydrolyzed to ferric hydroxide and ferrous hydroxide under alkali conditions.

The findings of Cao et al. (2008) were different from the above articles. They testified that both low (pH = 1.5) and high pH (pH = 11.0) were more favorable for oxidation of Lindane in Fe2+/PS system. The proposed reason is that Lindane is easily catalyzed hydrolysis by acid or base.

Anions concentration

Some inorganic anions, such as HCO3, NO3, NO2, SO42−, H2PO4 and Cl, in the water either have a positive or negative effect on sulfate radicals-based oxidation. This inorganic anion can either promote or inhibit the degradation of refractory organics in the Fe2+/PS system. The conclusions from some articles (Xu & Li 2010; Ghauch et al. 2013; Rao et al. 2014; Naim & Ghauch 2016; Bu et al. 2017; Amasha et al. 2018; Nie et al. 2018; Song et al. 2019; Wang et al. 2019c) indicated that NO3, SO42−, H2PO4 and HPO42− could inhibit degradation of refractory organics, while Cl, HCO3 could play both positive and negative effects on refractory organics removal. The inhibition effect of NO3, SO42−, H2PO4, HCO3 and Cl is attributed to the reactions occurring between inorganic anions and SO4•− (Equations (17)–(20)). H2PO4, NO3, Cl and HCO3 could react with SO4•−, then generate HPO4•−, NO3•−, Cl and CO3•2−, respectively. Because the redox potential of HPO4•−, NO3•−, Cl and CO3•2− are lower than that of SO4•−, the replacement of these free radicals with SO4•− leads to a decrease in the rate of degradation of refractory organics in Fe2+/PS system. Phosphate is a special case in PS systems activated by iron species. In fact phosphate is a good complexant for Fe2+ species to form FeH2PO42+ and FePO4 (Ghauch 2008) and this is why inhibition occurs. Unless done in Fe2+ free solution, phosphate can generate in heat activated PS medium some phosphate radicals as mentioned here (Ghauch & Tuqan 2012).

As for the effect of SO42− on refractory organics removal, the possible reason is due to the high ion strength that hinders the decomposition of PS (Rao et al. 2014). In addition, SO42− can reduce the reduction potential of half reaction of SO4•−, which also results in the reducing ability of SO4•− oxidation on refractory organics (Rao et al. 2014). One proposal for the positive effect of Cl on the organics degradation is that the formation of chlorine-containing radicals (Cl and Cl2•−) and free available chlorine (Cl2) through a series of chain reactions (Equations (21)–(24)) enhances the refractory organics removal (Rao et al. 2014; Wang et al. 2019c). Among these free radicals, Cl2•− as a stronger oxidizing property is major free radical when the Cl concentration reached 10.0 mM (Wang et al. 2019c). From the research of Song et al., 100 mM HCO3 enhanced TPhP degradation to 66.6% in Fe2+/PS process (Song et al. 2019). There are two possible reasons to explain the promotion of TPhP degradation by carbonate (Song et al. 2019). Firstly, HCO3 can easily react with electron-rich organic pollutants like TPhP. Alternatively, high concentration of HCO3 can stimulate the decomposition of PS into SO4•−, and promote the TPhP removal.
formula
(17)
formula
(18)
formula
(19)
formula
(20)
formula
(21)
formula
(22)
formula
(23)
formula
(24)

Chelating agents

Because too much or too little Fe2+ in solution has an adverse effect on activating PS to generate SO4•−, the chelating agents were used in Fe2+/PS system to avoid the drawback of alone utilization of Fe2+. Various chelating agents were added to aqueous solution in the form of acids or salts to chelate with Fe2+, such as hydroxylamine (HA), citric acid (CA), oxalic acid (OA), tartaric acid (TA), ethylenediaminetetraacetate (EDTA), (S,S) – ethylenediamine – N, N′ – disuccinate (EDDS), sodium thiosulfate (Na2S2O3), diethylene triamine pentaacetic acid (DTPA) and epigallocatechin-3-gallate (EGCG) (Zhou et al. 2013; Han et al. 2014, 2015a; Ji et al. 2014; Wu et al. 2015, 2016; Bu et al. 2017; Yu et al. 2018). Han et al. (2014), Wu et al. (2015) and Wu et al. (2016) concluded that HA is an effective chelating agent that could alleviate Fe3+ accumulation and accelerate Fe2+ regeneration (Equations (25)–(27)); the results shows an addition of HA can promote refractory organics removal rate.
formula
(25)
formula
(26)
formula
(27)
In the article by Ji et al. (2014), three kinds of chelating agents, CA, EDTA, and EDDS, were used to assist Fe2+ in activating PS. The study of their effects on antibiotics degradation indicates that CA and EDTA both have some promoting effect on sulfamethoxazole (SMX) degradation, while CA, EDTA, and EDDS did not show a significant impact on ciprofloxacin (CIP) degradation, which means that the effects of the chelating agents varies from different targeted contaminants. The article by Zhou et al. (2013) testified the addition of Na2S2O3 could facilitate the removal of diuron by about 25% compared with no Na2S2O3 added under the same conditions. This was due to Na2S2O3, acting as a buffer, balancing the amount of Fe3+ and Fe2+, as Equations (28)–(30).
formula
(28)
formula
(29)
formula
(30)
EGCG as green tea extract, was tested by Bu et al. (2017), and the result showed that EGCG could enhance the degradation rate of atrazine from 70% to 100% between pH 2.0 and 7.0. Owing to the strong chelating and reducing ability of EGCG, it accelerated the transformation rate from Fe3+ to Fe2+, as showed in Equations (31)–(33) (Bu et al.2017).
formula
(31)
formula
(32)
formula
(33)

From the above research results on chelating agents, we know that some chelating agents can promote the efficiency of Fe2+ activation reaction under certain conditions in the Fe2+/PS process, and these reaction conditions need to be further explored. In another point, the chelating agents used in the Fe2+/PS system should be considered not to cause secondary pollution to the solution and to be easily biodegraded.

ZERO-VALENT IRON-RELATED ACTIVATORS

Zero-valent iron (Fe0)

Although Fe2+ appears to be a good activator for PS, Fe2+/PS systems still have some limitations such as scavenging effect of excess Fe2+ on SO4•−, conversion of Fe2+ to iron hydroxide precipitation, and acidic pH of the solution. Fe0 has been widely studied as an alternative source of Fe2+ because it can gradually release ferrous iron into the water.

In a Fe0/PS system, Fe0 is corroded under aerobic (Equation (34)) or anaerobic (Equation (35)) conditions to form Fe2+; the formed Fe2+ further reacts with PS to generate SO4•−, which is consistent with Equation (1). Moreover, based on Equation (36), SO4•− could be generated straight through electron transformation from Fe0 to S2O82−. According to Equation (37), the recycling of Fe2+ on the surface of Fe0 can effectively avoid the accumulation of excess Fe2+ so that the amount of iron hydroxide reproduction can be reduced in the Fe0/PS system.
formula
(34)
formula
(35)
formula
(36)
formula
(37)

Table 3 summarizes the reaction conditions and degradation efficiency of some refractory organics using Fe0 as an activator for PS to generate SO4•−. It can be seen from Table 3 that most of the activation reactions in Fe0/PS system can achieve high refractory organics removal rates of 70–100% in the range of pH 3.0–7.0. Some papers (Oh et al. 2009; Deng et al. 2014; Pulicharla et al. 2018; Zhang et al. 2018b) compared the effects of Fe0 and Fe2+ acting as activators for PS on refractory organics removal and all results showed that Fe0 has a better treatment effect on refractory organics degradation than Fe2+ under the same conditions of the reaction. Except, the results of Gao et al. (2018) showed that Fe0 and Fe2+ both had the same removal effect on bisphenol A removal at neutral pH. Under most conditions, the activation effect of Fe0 on PS is still better than that of Fe2+ on PS. Similar to previously discussed Fe2+/PS system, the molar ratio of catalyst to oxidant is an important parameter in Fe0/PS system. Low Fe0:PS molar ratio will lead to poor degradation effect on refractory organics due to lack of sufficient active sites for PS activation and SO4•− generation, while high Fe0: PS molar ratio will result in less SO4•− production and unnecessary Fe0 consumption. According to previous analysis, we know the theoretically optimal ratio of Fe2+: PS was 1:1. However, Fe0: PS molar ratio needs to be higher than 1:1 in the bulk solution in order to ensure enough Fe0 that can consistently liberate Fe2+ into the water. Moreover, the ratio of catalyst to oxidant ultimately should be determined under specific reaction conditions.

Table 3

Reaction conditions and EC treatment effects in the Fe0/ PS system

Pollutant (Abbreviation)Parameters and degradation rateRemarksReference
Polyvinyl alcohol (PVA) [Fe0]/[PS] = 1:1,120 min, T = 20 °C, [PS] = 250 mg/L, PVA = 47.1–50.2 mg/L, 100% The optimal persulfate to Fe2+ or Fe0 molar ratio was found to be 1:1. Oh et al. (2009)  
Naphthalene (NAP) [NAP] = 10 mg/L, [PS] = 42 mM, [Fe0] = 1 g/L, T = 20 °C,pH = 5.0, 3 min, 99% The initial pH of 5.0 changed to around 2.3 in the NAP/PS/Fe0 systems after reaction. Liang & Guo (2010)  
2,4-dinitrotoluene (DNT) [PS] = 250 mg/L, [Fe0] = 1 g/L, [DNT] = 50 mg/L, t = 300 min, 91% As a persulfate-activating agent, Fe0 is more effective and longer-lasting than Fe2+ and potentially more suitable for environmental applications. Oh et al. (2010)  
4-chlorophenol [Fe0] = 0.20 g/L, [4-chlorophenol] = 0.156 mM; [PS] = 0.78 mM, no pH adjustment, t = 1 h, 88% The removal of 4-chlorophenol with Fe0-PS was 86% and 80% at pH 3.0 and 6.0, respectively. Zhao et al. (2010)  
p-chloroaniline (PCA) [PCA] = 0.05 mM, [PS] = 2.5 mM, [Fe0] = 0.70 g/L, pH = 6.8, t = 300 min, 64.55% The PCA degradation was higher under acidic conditions (pH 2.0 and 4.0) compared to alkaline conditions. Complete degradation of PCA was obtained by ZVI-activated persulfate at pH 4.0 at 12 min. Hussain et al. (2012)  
Sulfamethoxazole (SMX) [Fe0] = 17.85 mM, [PS] = 1.0 mM, [SMX] = 39.5 ìM, pH = 5.54, t = 2 h, 91%  Reactions with acid-washed MIPs (wFe/SMX) were more successful than those carried out with un-washed MIPs (nwFe/SMX). Ghauch et al. (2013)  
Acetaminophen (APAP) [Fe0]:[PS] = 1:1,180 min, 93.19% The Fe0/PS system was effective in a broader pH range from 3.0 to 8.5; heat could enhance the APAP degradation. Deng et al. (2014)  
Aniline [Fe0] = 0.4 g/L, [Aniline] = 0.05 mM, T = 25 °C, [PS] = 2.5 mM, pH = 7.0, t = 300 min, 85% The optimum pH for degradation of aniline was 4.0 in which 100% degradation efficiency was achieved after 10 min reaction. Hussain et al. (2014)  
Dibutyl phthalate (DBP) [DBP] = 5 mg/L; [PS] = 43 mg/L; [Fe0] = 0.3 g/L; T = 25 C, pH = 3.0, t = 120 min, 84% The half-lives for the oxidation of DBP in Fe0/PS were 30–176 min at pH 3.0–11.0. Li et al. (2014)  
Acid Orange 7 (AO7) [AO7] = 30 mg/L,[PS] = 0.3 g/L, [Fe0] = 0.5 g/L, pH = 5.8,20 min, 38%  Wang et al. (2014)  
Sulfadiazine (SD) [SD] = 20 mg/L, [PS] = 1.84 mM, [Fe0] = 0.92 mM, pH = 7.0, t = 1 h, 45.5%  Zou et al. (2014)  
1,1,1-trichloroethane (TCA) [PS] = 9.0 mM, [TCA] = 0.15 mM, [Fe0] = 2.08 g/L, T = 25 °C, t = 6 h, 90% There was no significant difference on degradation efficiency of TCA under pH 6–8. Gu et al. (2015)  
Acid orange 7 (AO7) [PS]:[AO7] = 30:1, [AO7] = 0.2 mM, [Fe0] = 0.5 g /L, pH = 3.8, T = 25 °C, 10 min, 88% AO7 removal The overall oxidation rate of AO7 was inhibited upon addition of NO3, NO2, SO42−, Cl, CO32−, HCO3, HPO42−, H2PO4 and EDTA, whereas ClO4, CH3COO and HA were found to accelerate AO7 decolorization rates. Li et al. (2015)  
Chloramphenicol (CAP) [CAP] = 0.05 mM, [Fe0] = 1 mM, [PS] = 1 mM, pH = 5.4,50 min, 95% The optimum molar ratio of CAP/ZVI/PS is 1:20:20. Nie et al. (2015)  
Sirius Red F3B (SRF3B) [SRF3B] = 25 mg/L, pH = 6.0, [PS] = 5 mM, [Fe0] = 0.5 g/L, 10 min, 95% decolorization 120 W/L US The PS/(Fe0) system shows a practically feasible process for SRF3B decolorization under conditions of initial pH <9.0, high PS and Fe0 dosages, and high temperatures. Weng et al. (2015
Para-Chlorophenol (PCP) [PS] = 15 mM, [Fe0] = 1 g/L, [PCP] = 50 mg/L, pH = 5.0, T = 25 °C, t = 120 min, 70% The degradation efficiency of PCP gradually decreases to 51.7% when the ZVI reuse time is increased. Ahmadpour & Yengejeh (2016)  
Triton X − 45 (TX − 45) [TX − 45] = 20 mg/L, [Fe0] = 1 g/L, [PS] = 2.5 mM, pH = 5.0, t = 60 min, 90% TX − 45 and TOC removals were most efficient under acidic and neutral pH valueswith addition of 2.5 mM PS. Temiz et al. (2016)  
Alachlor [Alachlor] = 5 mg/L,[Fe0] = 2.0 mM, [PS] = 1.0 mM, T = 30 °C, 60 min, 100% The reaction rate is significantly influenced by factors such as Fe0 dosage, PS dosage, initial pH, temperature, NOM, citrate and anions. Wang et al. (2016b)  
Direct Red 23 (DR23) [DR23] = 0.1 mM, [PS] = 5 mM, [Fe0] = 0.5 g/L, pH = 6,T = 25 °C, t = 15 min, 95% A promising decolorization was observed in the PS oxidation process activated with Fe0 aggregates. The decolorization efficiency in the PS/Fe0 system was further enhanced by US (PS/Fe0/US) or heat (PS/Fe0/55 °C). Weng & Tsai (2016)  
Orange G (OG) [OG] = 0.2 mM, [PS] = 2 mM, [Fe0] = 0.2 g/L, CuO = 0.1 g/L,T = 25 °C, pH = 5.0, t = 1 h, 50%  Wang et al. (2016a)  
Bentazon (BTZ) [Fe0] = 4.477 mM,[PS] = 0.262 mM, [BTZ] = 0.021 mM, pH ≤ 7, 100% Al3+, Cl, NO3 improved the treatment; NH4+, Ca2+, and Mg2+ did not significantly influence the BTZ removal; and Mn2+, Cu2+, CO32− HCO3, PO43−, HPO42−, H2PO4 inhibited the BTZ degradation. Wei et al. (2016)  
Reactive Green 19 (RG19) [RG19] = 0.01 mM, [PS] = 5 mM, [Fe0] = 1.0 g/L, pH = 6.0, T = 25 °C, 10 min, 99% sodium citrate, sodium EDTA, and [sodium oxalate] = 1 mM RG19 decolorization decreased from 99% to 7%, 21%, and 15% in presence of sodium citrate, sodium EDTA, and sodium oxalate, compared with control test (without chelating agent, 99%) within 10 min. Ding et al. (2017)  
p-nitrophenol (PNP) [PNP] = 500 mg/L, [Fe0] = 30 g/ L, [PS] = 12.5 mM, pH = 3.6, T = 25 °C, 3 min, 95%  Ji et al. (2017)  
1,4-dioxane (1,4-D) [Fe(0)]:[PS] = 1:1, [Fe0] = 67 mM, [PS] = 63 mM,[PS] :[dioxane] = 11:1, t = 10 min, 60% of dioxane removal Distributing the Fe0 as eight, one-eighth dosages, 95% dioxane removal was achieved; adding sodium hexametaphosphate (SHMP) to the persulfate candles can facilitate persulfate release. Kambhu et al. (2017)  
p-nitrophenol (PNP) [Fe0] = 1.3 g/L, [PS] = 6.7 mM, pH = 5.1, 20 min, 29.6%  Li. et al. (2017a) 
Erythromycin (ERY) [Fe0] = 22.4 mg L, [PS] = 0.1 mM; [ERY] = 1 mg L,T = 25 °C, initial pH, t = 240 min, 91.2% The ERY removal was more favorable under acidic and neutral condition than alkaline condition. Li et al. (2017b)  
2,4-dichlorophenol (2,4-DCP) [Fe0] = 1.5 mM, [2,4-DCP] = 4 mg/L, [PS] = 1.0 mM, pH = 7,15 min, 90% The Pre-Fe0/PS system induced a significant improvement in removal efficiency of 2,4-DCP, TOC and de-chlorination, especially at a higher initial pH condition. Li. et al. (2017c)  
Orange G (OG) [Fe0] = 4 mM, [PS] = 4 mM, [OG] = 200 mg/L, pH = 7, 180 min, 28.6%  Pan et al. (2017)  
Sulfadiazine (SDZ) [Fe0] = 1 mM, [PS] = 1 mM, [SDZ] = 20 mM, T = 25 °C, pH = 7,10 min, 80% Common aquatic materials including sulfate, nitrate, chloride, perchlorate, and HA all showed a negative effect on SDZ degradation by Fe0/PS following a trend of Cl < ClO4 < SO42− < NO3 <HCO3 < HA. Yang & Che (2017)  
Bisphenol A (BPA) [BPA] = 0.0877 mM, [PS] = 4.385 mM, [Fe0] = 0.0877 mM–0.877 mM pH = 7.0, T = 25 °C, t = 60 min, 85.17% − 100% The best BPA decomposition effects were obtained when [Fe2+]:[PS] = 1:2 and [Fe0]: [PS] = 1:2 are applied. Gao et al. (2018)  
Chlortetracycline (CTC) [Fe0] = 1,000 mM, 94% Whereas in heterogeneous reactions, slow generation and regeneration of Fe2+ ions obtained a removal of CTC above 90%. Pulicharla et al. (2018)  
Rhodamine B (RhB) [RhB] = 50 mg/ L, [PS] = 1.4 g/L, H2A/Fe0 = 1.0 g/ L, T = 25 °C, 60 min, 58% The optimal PS dosage is 0.3 g/L. Wang et al. (2018)  
Remazol Golden Yellow (RGY) [PS] = 5 × 10–3 M, [Fe0] = 0.5 g/L, [RGY] = 100 mg/L, pH = 6.0, t = 20 min, 98% The inhibitory effect of various inorganic salts on decolorization follows the sequence of Na2HPO4> NaHCO3 > NaClO4 > NaCl > NaNO3 > NaClO4 > no salt. Weng & Tao (2018)  
2,4,6-tricholoroanisole (TCA) [PS] = 0.24 mM, [Fe0] = 0.48 mM, pH = 7, [TCA] = 0.47ìM, temperature: T = 25 °C; t = 60 min, 18% CA/Fe0 molar ratio of 2:5, CA promoted the oxidation of TCA by Fe0/PS, but OA and EDTA retarded the reaction. Citric acid(CA),oxalic acid(OA)Fe0/PS had the best performance at pH 2.5. Zhang et al. (2018b)  
Pollutant (Abbreviation)Parameters and degradation rateRemarksReference
Polyvinyl alcohol (PVA) [Fe0]/[PS] = 1:1,120 min, T = 20 °C, [PS] = 250 mg/L, PVA = 47.1–50.2 mg/L, 100% The optimal persulfate to Fe2+ or Fe0 molar ratio was found to be 1:1. Oh et al. (2009)  
Naphthalene (NAP) [NAP] = 10 mg/L, [PS] = 42 mM, [Fe0] = 1 g/L, T = 20 °C,pH = 5.0, 3 min, 99% The initial pH of 5.0 changed to around 2.3 in the NAP/PS/Fe0 systems after reaction. Liang & Guo (2010)  
2,4-dinitrotoluene (DNT) [PS] = 250 mg/L, [Fe0] = 1 g/L, [DNT] = 50 mg/L, t = 300 min, 91% As a persulfate-activating agent, Fe0 is more effective and longer-lasting than Fe2+ and potentially more suitable for environmental applications. Oh et al. (2010)  
4-chlorophenol [Fe0] = 0.20 g/L, [4-chlorophenol] = 0.156 mM; [PS] = 0.78 mM, no pH adjustment, t = 1 h, 88% The removal of 4-chlorophenol with Fe0-PS was 86% and 80% at pH 3.0 and 6.0, respectively. Zhao et al. (2010)  
p-chloroaniline (PCA) [PCA] = 0.05 mM, [PS] = 2.5 mM, [Fe0] = 0.70 g/L, pH = 6.8, t = 300 min, 64.55% The PCA degradation was higher under acidic conditions (pH 2.0 and 4.0) compared to alkaline conditions. Complete degradation of PCA was obtained by ZVI-activated persulfate at pH 4.0 at 12 min. Hussain et al. (2012)  
Sulfamethoxazole (SMX) [Fe0] = 17.85 mM, [PS] = 1.0 mM, [SMX] = 39.5 ìM, pH = 5.54, t = 2 h, 91%  Reactions with acid-washed MIPs (wFe/SMX) were more successful than those carried out with un-washed MIPs (nwFe/SMX). Ghauch et al. (2013)  
Acetaminophen (APAP) [Fe0]:[PS] = 1:1,180 min, 93.19% The Fe0/PS system was effective in a broader pH range from 3.0 to 8.5; heat could enhance the APAP degradation. Deng et al. (2014)  
Aniline [Fe0] = 0.4 g/L, [Aniline] = 0.05 mM, T = 25 °C, [PS] = 2.5 mM, pH = 7.0, t = 300 min, 85% The optimum pH for degradation of aniline was 4.0 in which 100% degradation efficiency was achieved after 10 min reaction. Hussain et al. (2014)  
Dibutyl phthalate (DBP) [DBP] = 5 mg/L; [PS] = 43 mg/L; [Fe0] = 0.3 g/L; T = 25 C, pH = 3.0, t = 120 min, 84% The half-lives for the oxidation of DBP in Fe0/PS were 30–176 min at pH 3.0–11.0. Li et al. (2014)  
Acid Orange 7 (AO7) [AO7] = 30 mg/L,[PS] = 0.3 g/L, [Fe0] = 0.5 g/L, pH = 5.8,20 min, 38%  Wang et al. (2014)  
Sulfadiazine (SD) [SD] = 20 mg/L, [PS] = 1.84 mM, [Fe0] = 0.92 mM, pH = 7.0, t = 1 h, 45.5%  Zou et al. (2014)  
1,1,1-trichloroethane (TCA) [PS] = 9.0 mM, [TCA] = 0.15 mM, [Fe0] = 2.08 g/L, T = 25 °C, t = 6 h, 90% There was no significant difference on degradation efficiency of TCA under pH 6–8. Gu et al. (2015)  
Acid orange 7 (AO7) [PS]:[AO7] = 30:1, [AO7] = 0.2 mM, [Fe0] = 0.5 g /L, pH = 3.8, T = 25 °C, 10 min, 88% AO7 removal The overall oxidation rate of AO7 was inhibited upon addition of NO3, NO2, SO42−, Cl, CO32−, HCO3, HPO42−, H2PO4 and EDTA, whereas ClO4, CH3COO and HA were found to accelerate AO7 decolorization rates. Li et al. (2015)  
Chloramphenicol (CAP) [CAP] = 0.05 mM, [Fe0] = 1 mM, [PS] = 1 mM, pH = 5.4,50 min, 95% The optimum molar ratio of CAP/ZVI/PS is 1:20:20. Nie et al. (2015)  
Sirius Red F3B (SRF3B) [SRF3B] = 25 mg/L, pH = 6.0, [PS] = 5 mM, [Fe0] = 0.5 g/L, 10 min, 95% decolorization 120 W/L US The PS/(Fe0) system shows a practically feasible process for SRF3B decolorization under conditions of initial pH <9.0, high PS and Fe0 dosages, and high temperatures. Weng et al. (2015
Para-Chlorophenol (PCP) [PS] = 15 mM, [Fe0] = 1 g/L, [PCP] = 50 mg/L, pH = 5.0, T = 25 °C, t = 120 min, 70% The degradation efficiency of PCP gradually decreases to 51.7% when the ZVI reuse time is increased. Ahmadpour & Yengejeh (2016)  
Triton X − 45 (TX − 45) [TX − 45] = 20 mg/L, [Fe0] = 1 g/L, [PS] = 2.5 mM, pH = 5.0, t = 60 min, 90% TX − 45 and TOC removals were most efficient under acidic and neutral pH valueswith addition of 2.5 mM PS. Temiz et al. (2016)  
Alachlor [Alachlor] = 5 mg/L,[Fe0] = 2.0 mM, [PS] = 1.0 mM, T = 30 °C, 60 min, 100% The reaction rate is significantly influenced by factors such as Fe0 dosage, PS dosage, initial pH, temperature, NOM, citrate and anions. Wang et al. (2016b)  
Direct Red 23 (DR23) [DR23] = 0.1 mM, [PS] = 5 mM, [Fe0] = 0.5 g/L, pH = 6,T = 25 °C, t = 15 min, 95% A promising decolorization was observed in the PS oxidation process activated with Fe0 aggregates. The decolorization efficiency in the PS/Fe0 system was further enhanced by US (PS/Fe0/US) or heat (PS/Fe0/55 °C). Weng & Tsai (2016)  
Orange G (OG) [OG] = 0.2 mM, [PS] = 2 mM, [Fe0] = 0.2 g/L, CuO = 0.1 g/L,T = 25 °C, pH = 5.0, t = 1 h, 50%  Wang et al. (2016a)  
Bentazon (BTZ) [Fe0] = 4.477 mM,[PS] = 0.262 mM, [BTZ] = 0.021 mM, pH ≤ 7, 100% Al3+, Cl, NO3 improved the treatment; NH4+, Ca2+, and Mg2+ did not significantly influence the BTZ removal; and Mn2+, Cu2+, CO32− HCO3, PO43−, HPO42−, H2PO4 inhibited the BTZ degradation. Wei et al. (2016)  
Reactive Green 19 (RG19) [RG19] = 0.01 mM, [PS] = 5 mM, [Fe0] = 1.0 g/L, pH = 6.0, T = 25 °C, 10 min, 99% sodium citrate, sodium EDTA, and [sodium oxalate] = 1 mM RG19 decolorization decreased from 99% to 7%, 21%, and 15% in presence of sodium citrate, sodium EDTA, and sodium oxalate, compared with control test (without chelating agent, 99%) within 10 min. Ding et al. (2017)  
p-nitrophenol (PNP) [PNP] = 500 mg/L, [Fe0] = 30 g/ L, [PS] = 12.5 mM, pH = 3.6, T = 25 °C, 3 min, 95%  Ji et al. (2017)  
1,4-dioxane (1,4-D) [Fe(0)]:[PS] = 1:1, [Fe0] = 67 mM, [PS] = 63 mM,[PS] :[dioxane] = 11:1, t = 10 min, 60% of dioxane removal Distributing the Fe0 as eight, one-eighth dosages, 95% dioxane removal was achieved; adding sodium hexametaphosphate (SHMP) to the persulfate candles can facilitate persulfate release. Kambhu et al. (2017)  
p-nitrophenol (PNP) [Fe0] = 1.3 g/L, [PS] = 6.7 mM, pH = 5.1, 20 min, 29.6%  Li. et al. (2017a) 
Erythromycin (ERY) [Fe0] = 22.4 mg L, [PS] = 0.1 mM; [ERY] = 1 mg L,T = 25 °C, initial pH, t = 240 min, 91.2% The ERY removal was more favorable under acidic and neutral condition than alkaline condition. Li et al. (2017b)  
2,4-dichlorophenol (2,4-DCP) [Fe0] = 1.5 mM, [2,4-DCP] = 4 mg/L, [PS] = 1.0 mM, pH = 7,15 min, 90% The Pre-Fe0/PS system induced a significant improvement in removal efficiency of 2,4-DCP, TOC and de-chlorination, especially at a higher initial pH condition. Li. et al. (2017c)  
Orange G (OG) [Fe0] = 4 mM, [PS] = 4 mM, [OG] = 200 mg/L, pH = 7, 180 min, 28.6%  Pan et al. (2017)  
Sulfadiazine (SDZ) [Fe0] = 1 mM, [PS] = 1 mM, [SDZ] = 20 mM, T = 25 °C, pH = 7,10 min, 80% Common aquatic materials including sulfate, nitrate, chloride, perchlorate, and HA all showed a negative effect on SDZ degradation by Fe0/PS following a trend of Cl < ClO4 < SO42− < NO3 <HCO3 < HA. Yang & Che (2017)  
Bisphenol A (BPA) [BPA] = 0.0877 mM, [PS] = 4.385 mM, [Fe0] = 0.0877 mM–0.877 mM pH = 7.0, T = 25 °C, t = 60 min, 85.17% − 100% The best BPA decomposition effects were obtained when [Fe2+]:[PS] = 1:2 and [Fe0]: [PS] = 1:2 are applied. Gao et al. (2018)  
Chlortetracycline (CTC) [Fe0] = 1,000 mM, 94% Whereas in heterogeneous reactions, slow generation and regeneration of Fe2+ ions obtained a removal of CTC above 90%. Pulicharla et al. (2018)  
Rhodamine B (RhB) [RhB] = 50 mg/ L, [PS] = 1.4 g/L, H2A/Fe0 = 1.0 g/ L, T = 25 °C, 60 min, 58% The optimal PS dosage is 0.3 g/L. Wang et al. (2018)  
Remazol Golden Yellow (RGY) [PS] = 5 × 10–3 M, [Fe0] = 0.5 g/L, [RGY] = 100 mg/L, pH = 6.0, t = 20 min, 98% The inhibitory effect of various inorganic salts on decolorization follows the sequence of Na2HPO4> NaHCO3 > NaClO4 > NaCl > NaNO3 > NaClO4 > no salt. Weng & Tao (2018)  
2,4,6-tricholoroanisole (TCA) [PS] = 0.24 mM, [Fe0] = 0.48 mM, pH = 7, [TCA] = 0.47ìM, temperature: T = 25 °C; t = 60 min, 18% CA/Fe0 molar ratio of 2:5, CA promoted the oxidation of TCA by Fe0/PS, but OA and EDTA retarded the reaction. Citric acid(CA),oxalic acid(OA)Fe0/PS had the best performance at pH 2.5. Zhang et al. (2018b)  

Fe0 shows a good activated effect on PS. However, high concentrations of Fe0 and PS are required, which increases treatment cost. In order to improve efficiency of Fe0 on PS decomposition, researchers used two thin cylindrical neodymium-iron-boron permanent magnets to make Fe0 magnetic, and magnetized Fe0 (or pre-magnetization Fe0) was employed to activate PS for degrading refractory organics, such as orange (OG), 2,4-dichlorophenol (2,4-DCP) (Li et al. 2017c; Pan et al. 2017). It was found that in SEM micrographs, after pre-magnetization, the Fe0 surface was slightly corroded and deposition of tiny particles it was noted on the surface and the BET after pre-magnetization significantly increased from 0.15 m2/g (original Fe0) to 1.94 m2/g, which benefited the mass transfer of Fe2+ released from Fe0 (Li et al. 2017c). For these reasons, the degradation rate of refractory organics in the pre-magnetized Fe0/PS is also higher than Fe0/PS system (Li et al. 2017c; Pan et al. 2017).

Nano-sized zero-valent iron

To further improve the efficiency of Fe0/PS system, nFe0 has been used to activate PS for refractory organics removal because of high activity of nFe0 compared with Fe0. Better refractory organics removal efficiency has been found in nFe0/PS system than Fe2+/PS or Fe0/PS (Al-Shamsi & Thomson 2013; Zhu et al. 2016; Wang et al. 2017; Kim et al. 2018; Lin & Chen 2018). Lin & Chen (2018) used nFe0 for PS decomposition to degrade sulfamethazine (SMT), and the SMT removal rate of 93% was obtained in 5 min with addition of nFe0 of 56 mg/L and PS of 2 mM. Al-Shamsi & Thomson (2013) compared the degradation rate of TCE using three different iron activators of nFe0, Fe0 and Fe2+. The TCE degradation rate in 3 min by nFe0/PS was 1.11 × 10−4Mmin−1, which is higher than that of Fe0/PS (5.18 × 10−6Mmin−1) and Fe2+/PS (6.25 × 10−5Mmin−1). Zhu et al. (2016) also studied the effects of the three iron activators for DDT removal, and the result is similar to Wang et al. (2017). To sum up, it is proposed that more SO4•− were formed in nFe0/PS than Fe2+/PS or Fe0/PS because of large specific surface areas of nFe0. Although valence of the Fe0 and nFe0 is same, nFe0 has a different surface structure from Fe0. Generally, there is formed a cycle of Fe3+ and Fe2+ on the surface of Fe0. Al-Shamsi & Thomson (2013) used X-ray photoelectron spectroscopy analyzed the surface of nFe0; they found that there was a layer of FeOOH (ferric oxyhydroxideon) on the surface of fresh nFe0. Furthermore, a passivated iron sulfate layer was formed on the nFe0 surfaces as fresh nFe0 contacts with the PS. The mechanism of the FeOOH layer formation are presented on Equations (34)–(41).
formula
(38)
formula
(39)
formula
(40)
formula
(41)
The mechanism of formation of ferrous sulfate layer can be explained by some equations. The reaction between nFe0 and PS is as same as Fe0 and PS, and the products of the reactions in Fe0/PS system and Fe2+/PS system both include SO4•− and SO42− (Equation (36)). Some Fe2+ react with PS to form SO42− (Equation (1)), some Fe2+ scavenge SO4•− to generate SO42− (Equation (2)). Surplus SO42− can react with Fe0 or FeOOH to produce ferrous sulfate precipitated on the surface of nFe0 (Equations (42) and (43)). Although the layer of ferrous sulfate on the nFe0 surface will reduce reaction rate to some extent, the activation effect of nFe0 is still better than Fe0 and Fe2+ (Al-Shamsi & Thomson 2013).
formula
(42)
formula
(43)
Compared with the funding of Al-Shamsi & Thomson (2013), Kim et al. (2018) used XRD and SEM-EDS to analyze surface composition of nFe0 when the reaction was initiating in the Fe0/PS system, and it was found that both lepidocrocite (FeOOH) and schwertmannite (Fe8O8(OH)6SO4(s)) were formed on the nFe0 surface, and the reaction were as Equations (44) and (45).
formula
(44)
formula
(45)

As reaction time increase, FeOOH reacted with nFe0 to form magnetite (Fe3O4) through electron transfer of bulk conduction. Simultaneously, the internal Fe0 volume continued to shrink, and released Fe2+ and Fe3+. Kim et al. (2018) proposed a model of core-shell structure (Figure 1) including Fe0(s) core, the inner layer Fe3O4, and the outer layer Fe3+. Moreover, they found that nFe0 can release Fe2+ for PS at a very fast rate, which is faster than Fe0 under optimal conditions.

Figure 1

Schematic illustration of multi-layered core-shell structure of nFe0.

Figure 1

Schematic illustration of multi-layered core-shell structure of nFe0.

Nanosized zero-valent iron/support materials

Although nFe0 has a better effectiveness in activating PS for refractory organics degradation owing to its small particle size, large surface area and high reactivity, its application still faced some practical problems (Fu et al. 2014). Mueller et al. (2012), Fu et al. (2014) and Stefaniuk et al. (2016) all found that high surface energies and inherent magnetism of nFe0 lead to agglomerating of nFe0 particles, which limits the mobility of nFe0 and reduces electron transfer of nFe0. In addition, nFe0 is easily oxidized by the oxygen as it is exposed to the air, lack of stability and difficulty separating from the water is problematic in terms of practical application. In order to solve the weaknesses of nFe0 particles, some materials (such as biochar (BC), reduced graphene oxide (rGO), mesoporous carbon (MC) and montmorillonite) have been used to support nFe0.

Biochar (BC) is a low-cost carbonaceous adsorbent generated in an oxygen-limited or oxygen-free environment at high temperatures by pyrolysis of biomass (Yan et al. 2015; Xu et al. 2018b; Luo et al. 2019; Park et al. 2019; Wang et al. 2019a, 2019d). BC has a strong adsorption capacity due to its high surface area (SA)-to-volume ratio and porous properties (Wei et al. 2018; Wang et al. 2019b). Ahmed et al. (2016) and Ravikumar et al. (2018) found the surface of BC contains abundant oxygen-rich functional groups (such as hydroxyl (–OH) and carboxyl (–COOH)) which can quickly adsorb organic contaminants. Han et al. (2015b) testified because of these unique properties of these functional groups, BC has been used as mechanical supporting material for nFe0 to disperse and stabilize nFe0. In detail, BC has a large specific surface area and a large number of micropores, which can provide a support carbon skeleton for nFe0 to immobilize (Liu et al. 2018; Wang et al. 2019d). nFe0 is stabilized on surface of BC via adsorption, coordination, chelation and bridging (Jiang et al. 2019; Wang et al. 2019d).

Some researchers inferred the PS activation mechanisms in nFe0/BC/PS system, and the basic mechanisms are illustrated as follows (Yan et al. 2015; Xu et al. 2018b; Luo et al. 2019; Park et al. 2019). (1) As nFe0 disperses well on BC surface, Fe2+, gradually formed by oxidation of nFe0, activates PS to form SO4•−. (2) Nano-Fe0 directly activates PS to generate SO4•− and HO based on Equations (46) and (47). (3) The appearance of ferrous sulfate may induce PS decomposition. Al-Shamsi & Thomson (2013) proposed that ferrous sulfate might be formed on the nFe0 surface in two ways. One way is that Fe0 and SO42− in solution form FeSO4 (Equation (48)). In the other way, under anaerobic conditions, Fe3+ is formed via oxidation of Fe2+ (Equation (49)). Fe(OH)3 is formed as an intermediate (Equations (50) and (51)), then generates FeOOH (Equation (52)) which can react with SO42− in solution to generate FeSO4 (Equation (53)). (4) Except for BC surface adsorption, BC abundant oxygen functional groups of C-OOH and C-OH on the surface can directly catalyze PS to generate SO4•− for degradation of refractory organics (Equations (54) and (55)) (Yan et al. 2015; Zhou et al. 2015; Tan et al. 2016; Xu et al. 2018b; Zhang et al. 2018a).
formula
(46)
formula
(47)
formula
(48)
formula
(49)
formula
(50)
formula
(51)
formula
(52)
formula
(53)
formula
(54)
formula
(55)
Reduced graphene oxide (rGO), is a two-dimensional monolayer with abundant sp2-bonded carbon. It is a versatile matrix used as a support for nFe0 nanoparticles (Yang et al. 2015; Hao et al. 2018; Soubh et al. 2018; Xu et al. 2018a). Owing to the layered structure, rGO has a large specific surface area that provides abundant adsorption sites for the immobilization of nFe0. The mechanism of activating PS in nFe0/rGO is similar to that of nFe0/BC. Except for providing support for nFe0, the functional groups such as –OOH and –OH on rGO surface can also catalyze PS to release organic radicals and sulfate radicals, seen as Equations (56) and (57) (Duan et al. 2015; Wu et al. 2018).
formula
(56)
formula
(57)

Ahmad et al. (2015) had investigated different types of reactive oxygen species (ROS) generated in the nFe0/rGO/PS system. Because of their different reaction rate constant, tert-Butanol, NB, AN, and CT were employed as probe compounds to identify HO, HO, both HO and SO4•−, and O2•−, respectively. The result proves the presence of three free radicals in the nFe0/rGO/PS system. In acidic conditions, more SO4•− radicals were generated compared to HO radicals, while O2•− radicals prefer a basic condition. In order to avoid the leakage of nFe0 and promote the tight adhesion of nFe0 on RGO, polydopamine (PDA), as a polymeric stabilizer, has been studied to use as nFe0 coating. PDA has a specialized adhesive foot protein and can adhere on nFe0 surface to modify nFe0 due to abundant functional groups possessing strong negative charge with the ability to bind positively charged iron. A study by Gu et al. (2018) has also confirmed that the nFe0/PDA/rGO/PS system showed better performance on pollutants removal than the nFe0/rGO/PS system owing to the presence of PDA. Jiang et al. (2018) investigated removal efficiency of TC in different systems, the removal rate of TC in descending order is: nFe0/MC/PS > MC/PS > MC > nFe0/MC > nFe0/PS >nFe0 > PS. The highest removal efficiency (92.1%) was obtained in the nFe0/MC/PS system. This could attribute to a synergistic effect between Fe0 and MC. Besides, the reaction mechanism of the nFe0/MC/PS system is similar to that of the Fe0/MC/PS system.

In addition, montmorillonite (Mt) as naturally abundant mineral clay has been investigated for support nFe0 to activate PS in degrading refractory organics (Wu et al. 2019) because of its high specific surface area and cation exchange capacity. Cationic surfactant-modified Mt (organo-montmorillonite, OMt) is more hydrophobic and has showed better catalytic performance than Mt (Wu et al. 2019).

SOME IRON-CONTAINING COMPOUNDS

Some iron-containing compounds such as hematite (Fe2O3), magnetite (Fe3O4), and ferrous sulfide (FeS), are also explored to activate PS for refractory organics degradation.

Teel et al. (2011) evaluated the effects of several soil minerals including hematite on PS activation and found that hematite is capable of activating PS for removal organic pollutants in water. This effect is attributed to the octahedral sites of hematite containing Fe2+ that can react with PS to produce SO4•−, according to Equation (1). Figure 2 illustrates the activated mechanism in which PS is activated to generate sulfate radicals for degradation of organic pollutants. The mechanism of the activation of PS by hematite is similar to that by Fe2+, but due to the heterogeneous catalysis of hematite, catalytic reactions will occur both on the surface of the Fe2O3 and in the solution. Firstly, Fe2+ at some sites on the surface of the hematite will adsorb PS in aqueous solution, then react with PS to form SO4•−, SO42− and Fe3+. Fe3+ also reacts with PS to form Fe2+, which forms a cycle between Fe2+ and Fe3+ on the surface of hematite. The Fe2+ and Fe3+ leaking into the solution also undergo the same cyclic reaction as the surface. There is also the formation of SO4•− during this cycle. SO4•− formed on the surface of hematite or in the solution can capture and react contaminants to form by-products in water, which ultimately produce carbon dioxide and water. Free radical scavenging experiment in reference (Kermani et al. 2018) have shown that there is also HO that could be generated during oxidation of PS by hematite; HO also plays a role on degradation of refractory organics. The type of free radicals generated in PS/Fe2O3 system are determined by solution pH and the type of targeting contaminants. SO4•− are usually dominant under an acidic condition, while HO prefers a neutral and alkaline environment.

Figure 2

Diagram of the mechanism of Fe2O3 catalyzing persulfate.

Figure 2

Diagram of the mechanism of Fe2O3 catalyzing persulfate.

Magnetite nanoparticles (MNPs) have also been studied as a catalyst similar to hematite because they can work under mild reaction conditions and be easily separated and recovered. Magnetite is a magnetic particle, and nanometer magnetite can be synthesized by reverse co-precipitation among FeCl3•6H2O + FeSO4•4H2O + NH3•H2O. Literature (Leng et al. 2013, 2014; Sun et al. 2016; Zhu et al. 2018) demonstrated that treatment effects, reaction conditions and mechanisms of different types of synthetic Fe3O4 and its complexes with other materials (biochar, NrGO) to activate PS for effectively degradation of refractory organics. Magnetite contains Fe2+, which can stimulate decomposition of PS to generate free radicals. However, the Fe2+ may have not enough magnetite due to the slow transformation between Fe2+ and Fe3+ cycles. As a result, modifying surfaces of Fe3O4 MNPs and improving the redox cycle of Fe2+/Fe3+ is a very important research issue (Sun et al. 2016). Some studies investigated polyhydroquinone-coated Fe3O4, ascorbic acid modified Fe3O4, β-CD immobilized Fe3O4 as catalysts for the activation of PS (Leng et al. 2013, 2014; Sun et al. 2016; Zhu et al. 2018). According to Sun et al. (2016), ascorbic acid (H2A, also called vitamin C), a redox mediator, used to modify Fe3O4 MNPs could enhance the ability of activating PS. Under same conditions, the effect of H2A/Fe3O4 on PS activation was greater than that of Fe3O4 alone on PS activation. Under H2A/Fe3O4 activation, a removal rate (98.5%) of 2,4-DCP can be obtained, while only 35.1% degradation of 2,4-DCP was achieved in Fe3O4/PS system, which was testified that the introduction of H2A in Fe3O4/PS can effectively activate PS for targeting pollutants removal. Nowack (2002) and Dong et al. (2017) all testified that H2A is an effective chelating and reducing agent that can enhance iron ions released from minerals. The primary mechanism of improved refractory organics degradation in the H2A/Fe3O4/PS process is illustrated in Figure 3. First, H2A, as a chelating agent, attached on surface of Fe3O4 particles combines with Fe2+ to form chelated iron of Fe2+–H2A through the ligand exchange reaction. Then, Fe2+ enters into aqueous phase initiating a reaction that catalyzes PS to generate SO4•−. In addition, H2A can enhance the performance of reduction of Fe3+ to Fe2+ on the Fe3O4 nanoparticle surface and promote more release of Fe2+ from Fe3O4 into solution.

Figure 3

Primary mechanism occurred in the H2A/ Fe3O4/PS process.

Figure 3

Primary mechanism occurred in the H2A/ Fe3O4/PS process.

Compared with homogenous Fe2+/PS system, heterogeneous FeS/PS system also shows a good performance on refractory organics removal (Pu et al. 2014; Yuan et al. 2015; Zhang et al. 2017; Fan et al. 2018). Yuan et al. (2015) found that the application ranges of pH for p-chloroaniline (PCA) fast degradation was from 3.0 to 7.0, while the optimum pH range was from 5.0 to 7.0 for complete degradation of PCA at 150 min according the research of Fan et al. (2018). The PS activation mechanism was affected by the initial solution pH. The mechanism of FeS activating PS is shown in Figure 4. As a solid particle like Fe0, FeS could continuously release dissolved Fe2+ (Yuan et al. 2015) which would activate PS to produce SO4•−. In addition, as an electron donor, S2+ was able to reduce Fe3+ to Fe2+ (Fan et al. 2018). Sustainable release of Fe2+ and recycling of Fe3+ by FeS during reaction could avoid surplus Fe2+ consumption and Fe3+ accumulation. Based on this mechanism, when the initial solution pH was 9.0–11.0, the dissolved Fe2+ from FeS was not enough to activate PS and relatively low PCA removal rate was acquired.

Figure 4

Schematic diagram of the mechanism of FeS activating PS.

Figure 4

Schematic diagram of the mechanism of FeS activating PS.

ECONOMIC EVALUATION

Regarding the methods for iron-based materials to activate PS to degrade organics, we are not only concerned with the efficiency, but also with the cost. Since these methods have not yet been applied in engineering, the cost of chemical agents in small experiments based on different iron materials is calculated and analyzed according to the consumption and price of chemical agents. Table 4 shows the unit price of iron-based materials found on the website of Sigma. Table 5 lists the experimental data such as oxide and catalyst concentration and removal efficiency in several representative small experiments, and also shows the cost of processing 1 L of water after calculation (in US dollars). The cost of each process is the sum of the cost of the catalyst and the oxidizing agent used to achieve an organic removal rate of more than 95%. According to the calculation results, the order of chemical costs from high to low is Fe3O4/PS, nFe0/PS, FeSO4·7H2O/PS, FeS/PS, Fe2O3/PS, and Fe0/PS. Since the price of the analytical reagent is used as a unit price, the calculation cost is high, and the cost will be reduced in actual engineering. In addition, the difference in the degree of difficult degradation of pollutants also determines that the same system will generate different treatment costs. These data are for reference only for small experimental chemistry costs.

Table 4

Unit price of iron-based materials

Molecular formulaUnit price USD/kgMolecular weightSizeCAS number
FeSO4·7H2166 278.01  7782-63-0 
Na2S2O8 43.2 238.1  7775-27-1 
K2S2O8 170.4 270.32  7727-21-1 
Fe0 92 55.85 325 mesh 7439-89-6 
Nano-Fe0 7,760 55.85 60–80 nm 7439-89-6 
Fe2O3 66 159.69 <5 μm 1309-37-1 
Fe3O4 566 231.53 <5 μm 1317-61-9 
FeS 264.4 87.91  1317-37-9 
Molecular formulaUnit price USD/kgMolecular weightSizeCAS number
FeSO4·7H2166 278.01  7782-63-0 
Na2S2O8 43.2 238.1  7775-27-1 
K2S2O8 170.4 270.32  7727-21-1 
Fe0 92 55.85 325 mesh 7439-89-6 
Nano-Fe0 7,760 55.85 60–80 nm 7439-89-6 
Fe2O3 66 159.69 <5 μm 1309-37-1 
Fe3O4 566 231.53 <5 μm 1317-61-9 
FeS 264.4 87.91  1317-37-9 
Table 5

The parameters and costs of the iron-based processes

Catalytic systemParametersCosts (USD/L)Reference
Fe2+/PS [Na2S2O8] = 4.385 mM, [Fe2+] = 2.1925 mM, [BPA] = 0.0876 mM, pH = 7.0, 100% 0.146 Gao et al. (2018)  
Fe0/PS [Na2S2O8] = 4.385 mM, [Fe0] = 2.1925 mM, [BPA] = 0.0876 mM, pH = 7.0, 100% 0.056 Gao et al. (2018)  
nFe0/PS [Na2S2O8] = 1 mM,[nFe0] = 28 mg/L, [SMT] = 10 mg/L, pH = 5.3, 100% 0.227 Lin & Chen (2018)  
Fe2O3/PS [Na2S2O8] = 10 mM,[Fe2O3] = 0.1 g/L, [OG] = 0.1 mM, pH = 5.0, 95.0% 0.110 Ike et al. (2018)  
Fe3O4/PS [K2S2O8] = 1.2 mM,[Fe3O4] = 2.4 mM, [SMM] = 0.06 mM, pH = 6.4, 96.0% 0.369 Yan et al. (2011)  
FeS/PS [Na2S2O8] = 4 mM,[FeS] = 0.35 g/L, [PCA] = 0.2 mM, pH = 3.0, 97.3% 0.133 Fan et al. (2018)  
Catalytic systemParametersCosts (USD/L)Reference
Fe2+/PS [Na2S2O8] = 4.385 mM, [Fe2+] = 2.1925 mM, [BPA] = 0.0876 mM, pH = 7.0, 100% 0.146 Gao et al. (2018)  
Fe0/PS [Na2S2O8] = 4.385 mM, [Fe0] = 2.1925 mM, [BPA] = 0.0876 mM, pH = 7.0, 100% 0.056 Gao et al. (2018)  
nFe0/PS [Na2S2O8] = 1 mM,[nFe0] = 28 mg/L, [SMT] = 10 mg/L, pH = 5.3, 100% 0.227 Lin & Chen (2018)  
Fe2O3/PS [Na2S2O8] = 10 mM,[Fe2O3] = 0.1 g/L, [OG] = 0.1 mM, pH = 5.0, 95.0% 0.110 Ike et al. (2018)  
Fe3O4/PS [K2S2O8] = 1.2 mM,[Fe3O4] = 2.4 mM, [SMM] = 0.06 mM, pH = 6.4, 96.0% 0.369 Yan et al. (2011)  
FeS/PS [Na2S2O8] = 4 mM,[FeS] = 0.35 g/L, [PCA] = 0.2 mM, pH = 3.0, 97.3% 0.133 Fan et al. (2018)  

CONCLUSION AND PROSPECTS

In summary, a state-of-the-art review is presented on mechanisms, reaction conditions, influencing factors, and treatment effects of iron-based materials for refractory organics degradation via PS activation. The results of recently published articles affirm that activated-PS by iron-based materials has promising potentials for destruction of refractory organics in water and 100% degradation can be accomplished for some refractory organics under optimized conditions. However, challenges still remain in the improvement of refractory organics degradation rate in actual polluted water treatment using iron-based materials/PS processes, and the practical engineering application of these processes with low costs is also highly anticipated.

The activation performance of iron-based materials can be influenced by the properties and concentrations of activators, pollutants, oxidants, and water quality component along with the reaction conditions (such as pH, temperature, reaction time). Optimization of the experimental conditions is a cornerstone in achieving best removal. Deep research on environmentally friendly iron-chelators will help to increase Fe0 functional lifespan and improve removal rate of refractory organics in Fe0/PS system. Without a doubt, nFe0 shows the most promise for PS activation if the problems of nFe0 leakage and accumulation can be properly solved. BC and rGO as effective supports for nFe0 can reduce nFe0 leakage, but the increased cost should be taken into account in future engineering applications. Moreover, it is also necessary to optimize the nFe0/BC/PS or nFe0/rGO/PS process from the perspective of reaction kinetics and other reaction conditions. Other iron-containing materials (Fe2O3, Fe3O4, FeS, etc.) also are used in the catalysis activation mechanism of Fe2+ to activate PS. Therefore, the key issue is to identify the factors that affect the release of Fe2+ and maintain a dynamic balance between Fe2+ and Fe3+ when using these materials as activators. Novel reactor for the iron-based materials/PS system based on batch or continuous flow need to designed, and more pilot experiments have yet to be implemented.

In a word, because of the low cost, availability, and environmental friendliness of iron, iron-based materials activation method for PS is expected to have a better prospect with further research in this field.

ACKNOWLEDGEMENT

The research group acknowledges the financial support provided by the Natural Sciences and Engineering Research Council (NSERC) of Canada through the CREATE, Discovery and CRD programs and Queen's University. Pascale Champagne is thankful to the Canada Research Chairs Program. Yanjiao Gao thanks the China Scholarship Council for funding a visiting scholar (201808210024). Tiehong Song expresses gratitude for the funding (JJKH20200288KJ) of Jilin Province Education Department of China.

REFERENCES

REFERENCES
Ahmad
A.
Gu
X.
Li
L.
Lu
S.
Xu
Y.
Guo
X.
2015
Effects of pH and anions on the generation of reactive oxygen species (ROS) in nZVI-rGO-activated persulfate system
.
Water, Air, & Soil Pollution
226
(
11
),
369
.
https://doi.org/10.1007/s11270-015-2635-8
.
Ahmadpour
E.
Yengejeh
R. J.
2016
Application of a zero-valente iron-per sulfate system to treat petrochemical wastewater with high-total dissolved solids containing para-chlorophenol
.
Jundishapur Journal of Health Sciences
8
(
2
).
https://doi.org/10.17795/jjhs-35108
.
Ahmed
M. B.
Zhou
J. L.
Ngo
H. H.
Guo
W.
Chen
M.
2016
Progress in the preparation and application of modified biochar for improved contaminant removal from water and wastewater
.
Bioresource Technology
214
,
836
851
.
https://doi.org/10.1016/j.biortech.2016.05.057
.
Al-Shamsi
M. A.
Thomson
N. R.
2013
Treatment of organic compounds by activated persulfate using nanoscale zerovalent iron
.
Industrial & Engineering Chemistry Research
52
(
38
),
13564
13571
.
https://doi.org/10.1021/ie400387p
.
Amasha
M.
Baalbaki
A.
Ghauch
A.
2018
A comparative study of the common persulfate activation techniques for the complete degradation of an NSAID: the case of ketoprofen
.
Chemical Engineering Journal
350
,
395
410
.
https://doi.org/10.1016/j.cej.2018.05.118
.
Babuponnusami
A.
Muthukumar
K.
2011
Degradation of phenol in aqueous solution by Fenton, sono-Fenton and sono-photo-Fenton methods
.
Clean–Soil, Air, Water
39
(
2
),
142
147
.
https://doi.org/10.1002/clen.201000072
.
Brienza
M.
Katsoyiannis
I.
2017
Sulfate radical technologies as tertiary treatment for the removal of emerging contaminants from wastewater
.
Sustainability
9
(
9
),
1604
.
https://doi.org/10.3390/su9091604
.
Bu
L.
Shi
Z.
Zhou
S.
2016
Modeling of Fe(II)-activated persulfate oxidation using atrazine as a target contaminant
.
Separation and Purification Technology
169
,
59
65
.
https://doi.org/10.1016/j.seppur.2016.05.037
.
Bu
L.
Bi
C.
Shi
Z.
Zhou
S.
2017
Significant enhancement on ferrous/persulfate oxidation with epigallocatechin-3-gallate: simultaneous chelating and reducing
.
Chemical Engineering Journal
321
,
642
650
.
https://doi.org/10.1016/j.cej.2017.04.001
.
Cao
J.
Zhang
W. X.
Brown
D. G.
Sethi
D.
2008
Oxidation of lindane with Fe(II)-activated sodium persulfate
.
Environmental Engineering Science
25
(
2
),
221
228
.
https://doi.org/10.1089/ees.2006.0244
.
Caré
S.
Crane
R.
Calabrò
P. S.
Ghauch
A.
Temgoua
E.
Noubactep
C.
2013
Modeling the permeability loss of metallic iron water filtration systems
.
Clean–Soil, Air, Water
41
(
3
),
275
282
.
https://doi.org/10.1002/clen.201200167
.
Chen
J.
Feng
J.
Lu
S.
Shen
Z.
Du
Y.
Peng
L.
Nian
P.
Yuan
S.
Zhang
A.
2018
Non-thermal plasma and Fe2+ activated persulfate ignited degradation of aqueous crystal violet: degradation mechanism and artificial neural network modeling
.
Separation and Purification Technology
191
,
75
85
.
https://doi.org/10.1016/j.seppur.2017.09.016
.
Cheng
X.
Guo
H.
Zhang
Y.
Wu
X.
Liu
Y.
2017
Non-photochemical production of singlet oxygen via activation of persulfate by carbon nanotubes
.
Water Research
113
,
80
88
.
https://doi.org/10.1016/j.watres.2017.02.016
.
Csay
T.
Homlok
R.
Illes
E.
Takacs
E.
Wojnarovits
L.
2014
The chemical background of advanced oxidation processes
.
Israel Journal of Chemistry
54
(
3
),
233
241
.
https://doi.org/10.1002/ijch.201300077
.
Darsinou
B.
Frontistis
Z.
Antonopoulou
M.
Konstantinou
I.
Mantzavinos
D.
2015
Sono-activated persulfate oxidation of bisphenol A: kinetics, pathways and the controversial role of temperature
.
Chemical Engineering Journal
280
,
623
633
.
https://doi.org/10.1016/j.cej.2015.06.061
.
Deng
Y.
Zhao
R.
2015
Advanced oxidation processes (AOPs) in wastewater treatment
.
Current Pollution Reports
1
(
3
),
167
176
.
https://doi.org/10.1007/s40726-015-0015-z
.
Deng
J.
Shao
Y.
Gao
N.
Deng
Y.
Tan
C.
Zhou
S.
2014
Zero-valent iron/persulfate (Fe0/PS) oxidation acetaminophen in water
.
International Journal of Environmental Science and Technology
11
(
4
),
881
890
.
https://doi.org/10.1007/s13762-013-0284-2
.
Ding
F.
Chen
H.
Zhang
S.
Zhao
T.
Liu
N.
2017
Effect of chelating agents on Reactive Green 19 decolorization through Fe0-activated persulfate oxidation process
.
Journal of Environmental Management
200
,
325
334
.
https://doi.org/10.1016/j.jenvman.2017.05.089
.
Dong
H.
Qiang
Z.
Hu
J.
Sans
C.
2017
Accelerated degradation of iopamidol in iron activated persulfate systems: roles of complexing agents
.
Chemical Engineering Journal
316
,
288
295
.
https://doi.org/10.1016/j.cej.2017.01.099
.
Duan
X.
Sun
H.
Kang
J.
Wang
Y.
Indrawirawan
S.
Wang
S.
2015
Insights into heterogeneous catalysis of persulfate activation on dimensional-structured nanocarbons
.
Acs Catalysis
5
(
8
),
4629
4636
.
https://doi.org/10.1021/acscatal.5b00774
.
Epold
I.
Trapido
M.
Dulova
N.
2015
Degradation of levofloxacin in aqueous solutions by Fenton, ferrous ion-activated persulfate and combined Fenton/persulfate systems
.
Chemical Engineering Journal
279
,
452
462
.
https://doi.org/10.1016/j.cej.2015.05.054
.
Fast
S. A.
Gude
V. G.
Truax
D. D.
Martin
J.
Magbanua
B. S.
2017
A critical evaluation of advanced oxidation processes for emerging contaminants removal
.
Environmental Processes
4
(
1
),
283
302
.
https://doi.org/10.1007/s40710-017-0207-1
.
Fu
F.
Dionysiou
D. D.
Liu
H.
2014
The use of zero-valent iron for groundwater remediation and wastewater treatment: a review
.
Journal of Hazardous Materials
267
,
194
205
.
https://doi.org/10.1016/j.jhazmat.2013.12.062
.
Gao
F.
Li
Y.
Xiang
B.
2018
Degradation of bisphenol A through transition metals activating persulfate process
.
Ecotoxicology and Environmental Safety
158
,
239
247
.
https://doi.org/10.1016/j.ecoenv.2018.03.035
.
Ghatak
H. R.
2014
Advanced oxidation processes for the treatment of biorecalcitrant organics in wastewater
.
Critical Reviews in Environmental Science and Technology
44
(
11
),
1167
1219
.
https://doi.org/10.1080/10643389.2013.763581
.
Ghauch
A.
2008
Rapid removal of flutriafol in water by zero-valent iron powder
.
Chemosphere
71
(
5
),
816
826
.
https://doi.org/10.1016/j.chemosphere.2007.11.057
.
Ghauch
A.
2015
Iron-based metallic systems: an excellent choice for sustainable water treatment
.
FOG-Freiberg Online Geoscience
38
,
1
80
.
Ghauch
A.
Tuqan
A.
2008
Catalytic degradation of chlorothalonil in water using bimetallic iron-based systems
.
Chemosphere
73
(
5
),
751
759
.
https://doi.org/10.1016/j.chemosphere.2008.06.035
.
Ghauch
A.
Tuqan
A.
2009
Reductive destruction and decontamination of aqueous solutions of chlorinated antimicrobial agent using bimetallic systems
.
Journal of Hazardous Materials
164
(
2–3
),
665
674
.
https://doi.org/10.1016/j.jhazmat.2008.08.048
.
Ghauch
A.
Tuqan
A. M.
2012
Oxidation of bisoprolol in heated persulfate/H2O systems: kinetics and products
.
Chemical Engineering Journal
183
,
162
171
.
https://doi.org/10.1016/j.cej.2011.12.048
.
Ghauch
A.
Tuqan
A.
Assi
H. A.
2009
Antibiotic removal from water: elimination of amoxicillin and ampicillin by microscale and nanoscale iron particles
.
Environmental Pollution
157
(
5
),
1626
1635
.
https://doi.org/10.1016/j.envpol.2008.12.024
.
Ghauch
A.
Assi
H. A.
Bdeir
S.
2010a
Aqueous removal of diclofenac by plated elemental iron: bimetallic systems
.
Journal of Hazardous Materials
182
(
1–3
),
64
74
.
https://doi.org/10.1016/j.jhazmat.2010.05.139
.
Ghauch
A.
Assi
H. A.
Tuqan
A.
2010b
Investigating the mechanism of clofibric acid removal in Fe0/H2O systems
.
Journal of Hazardous Materials
176
(
1–3
),
48
55
.
https://doi.org/10.1016/j.jhazmat.2009.10.125
.
Ghauch
A.
Assi
H. A.
Baydoun
H.
Tuqan
A. M.
Bejjani
A.
2011a
Fe0-based trimetallic systems for the removal of aqueous diclofenac: mechanism and kinetics
.
Chemical Engineering Journal
172
(
2–3
),
1033
1044
.
https://doi.org/10.1016/j.cej.2011.07.020
.
Ghauch
A.
Baydoun
H.
Dermesropian
P.
2011b
Degradation of aqueous carbamazepine in ultrasonic/Fe0/H2O2 systems
.
Chemical Engineering Journal
172
(
1
),
18
27
.
https://doi.org/10.1016/j.cej.2011.04.002
.
Ghauch
A.
Ayoub
G.
Naim
S.
2013
Degradation of sulfamethoxazole by persulfate assisted micrometric Fe0 in aqueous solution
.
Chemical Engineering Journal
228
,
1168
1181
.
https://doi.org/10.1016/j.cej.2013.05.045
.
Graça
C. A.
de Velosa
A. C.
Teixeira
A. C. S.
2017
Amicarbazone degradation by UVA-activated persulfate in the presence of hydrogen peroxide or Fe2+
.
Catalysis Today
280
,
80
85
.
https://doi.org/10.1016/j.cattod.2016.06.044
.
Gu
X.
Lu
S.
Guo
X.
Sima
J.
Qiu
Z.
Sui
Q.
2015
Oxidation and reduction performance of 1,1,1-trichloroethane in aqueous solution by means of a combination of persulfate and zero-valent iron
.
RSC Advances
5
(
75
),
60849
60856
.
https://doi.org/10.1039/C5RA07655B
.
Gu
M.
Sui
Q.
Farooq
U.
Zhang
X.
Qiu
Z.
Lyu
S.
2018
Degradation of phenanthrene in sulfate radical based oxidative environment by nZVI-PDA functionalized rGO catalyst
.
Chemical Engineering Journal
354
,
541
552
.
https://doi.org/10.1016/j.cej.2018.08.039
.
Guerra-Rodríguez
S.
Rodríguez
E.
Singh
D. N.
Rodríguez-Chueca
J.
2018
Assessment of sulfate radical-based advanced oxidation processes for water and wastewater treatment: a review
.
Water
10
(
12
),
1828
.
https://doi.org/10.3390/w10121828
.
Guo
Y.
Zhou
J.
Lou
X.
Liu
R.
Xiao
D.
Fang
C.
Wang
Z.
Liu
J.
2014
Enhanced degradation of Tetrabromobisphenol A in water by a UV/base/persulfate system: kinetics and intermediates
.
Chemical Engineering Journal
254
,
538
544
.
https://doi.org/10.1016/j.cej.2014.05.143
.
Han
D.
Wan
J.
Ma
Y.
Wang
Y.
Huang
M.
Chen
Y.
Li
D.
Guan
Z.
Li
Y.
2014
Enhanced decolorization of Orange G in a Fe(II)-EDDS activated persulfate process by accelerating the regeneration of ferrous iron with hydroxylamine
.
Chemical Engineering Journal
256
,
316
323
.
https://doi.org/10.1016/j.cej.2014.06.006
.
Han
D.
Wan
J.
Ma
Y.
Wang
Y.
Li
Y.
Li
D.
Guan
Z.
2015a
New insights into the role of organic chelating agents in Fe(II) activated persulfate processes
.
Chemical Engineering Journal
269
,
425
433
.
https://doi.org/10.1016/j.cej.2015.01.106
.
Han
L.
Xue
S.
Zhao
S.
Yan
J.
Qian
L.
Chen
M.
2015b
Biochar supported nanoscale iron particles for the efficient removal of methyl orange dye in aqueous solutions
.
PLoS One
10
(
7
).
https://doi.org/10.1371/journal.pone.0132067
.
Hao
J.
Ji
L.
Li
C.
Hu
C.
Wu
K.
2018
Rapid, efficient and economic removal of organic dyes and heavy metals from wastewater by zinc-induced in-situ reduction and precipitation of graphene oxide
.
Journal of the Taiwan Institute of Chemical Engineers
88
,
137
145
.
https://doi.org/10.1016/j.jtice.2018.03.045
.
Hussain
I.
Zhang
Y.
Huang
S.
Du
X.
2012
Degradation of p-chloroaniline by persulfate activated with zero-valent iron
.
Chemical Engineering Journal
203
,
269
276
.
https://doi.org/10.1016/j.cej.2012.06.120
.
Hussain
I.
Zhang
Y.
Huang
S.
2014
Degradation of aniline with zero-valent iron as an activator of persulfate in aqueous solution
.
RSC Advances
4
(
7
),
3502
3511
.
https://doi.org/10.1039/C3RA43364A
.
Ike
I. A.
Linden
K. G.
Orbell
J. D.
Duke
M.
2018
Critical review of the science and sustainability of persulphate advanced oxidation processes
.
Chemical Engineering Journal
338
,
651
669
.
https://doi.org/10.1016/j.cej.2018.01.034
.
Ioan
I.
Wilson
S.
Lundanes
E.
Neculai
A.
2007
Comparison of Fenton and sono-Fenton bisphenol A degradation
.
Journal of Hazardous Materials
142
(
1–2
),
559
563
.
https://doi.org/10.1016/j.jhazmat.2006.08.015
.
Ji
Y.
Ferronato
C.
Salvador
A.
Yang
X.
Chovelon
J. M.
2014
Degradation of ciprofloxacin and sulfamethoxazole by ferrous-activated persulfate: implications for remediation of groundwater contaminated by antibiotics
.
Science of the Total Environment
472
,
800
808
.
https://doi.org/10.1016/j.scitotenv.2013.11.008
.
Ji
Y.
Dong
C.
Kong
D.
Lu
J.
Zhou
Q.
2015
Heat-activated persulfate oxidation of atrazine: implications for remediation of groundwater contaminated by herbicides
.
Chemical Engineering Journal
263
,
45
54
.
https://doi.org/10.1016/j.cej.2014.10.097
.
Jiang
X.
Wu
Y.
Wang
P.
Li
H.
Dong
W.
2013
Degradation of bisphenol A in aqueous solution by persulfate activated with ferrous ion
.
Environmental Science and Pollution Research
20
(
7
),
4947
4953
.
https://doi.org/10.1007/s11356-013-1468-5
.
Jiang
X.
Guo
Y.
Zhang
L.
Jiang
W.
Xie
R.
2018
Catalytic degradation of tetracycline hydrochloride by persulfate activated with nano Fe0 immobilized mesoporous carbon
.
Chemical Engineering Journal
341
,
392
401
.
https://doi.org/10.1016/j.cej.2018.02.034
.
Jiang
S. F.
Ling
L. L.
Chen
W. J.
Liu
W. J.
Li
D. C.
Jiang
H.
2019
High efficient removal of bisphenol A in a peroxymonosulfate/iron functionalized biochar system: mechanistic elucidation and quantification of the contributors
.
Chemical Engineering Journal
359
,
572
583
.
https://doi.org/10.1016/j.cej.2018.11.124
.
Kambhu
A.
Gren
M.
Tang
W.
Comfort
S.
Harris
C. E.
2017
Remediating 1, 4-dioxane-contaminated water with slow-release persulfate and zerovalent iron
.
Chemosphere
175
,
170
177
.
https://doi.org/10.1016/j.chemosphere.2017.02.044
.
Kang
J.
Duan
X.
Wang
C.
Sun
H.
Tan
X.
Tade
M. O.
Wang
S.
2018
Nitrogen-doped bamboo-like carbon nanotubes with Ni encapsulation for persulfate activation to remove emerging contaminants with excellent catalytic stability
.
Chemical Engineering Journal
332
,
398
408
.
https://doi.org/10.1016/j.cej.2017.09.102
.
Kermani
M.
Mohammadi
F.
Kakavandi
B.
Esrafili
A.
Rostamifasih
Z.
2018
Simultaneous catalytic degradation of 2, 4-D and MCPA herbicides using sulfate radical-based heterogeneous oxidation over persulfate activated by natural hematite (α-Fe2O3/PS)
.
Journal of Physics and Chemistry of Solids
117
,
49
59
.
https://doi.org/10.1016/j.jpcs.2018.02.009
.
Kim
C.
Ahn
J. Y.
Kim
T. Y.
Shin
W. S.
Hwang
I.
2018
Activation of persulfate by nanosized zero-valent iron (NZVI): mechanisms and transformation products of NZVI
.
Environmental Science & Technology
52
(
6
),
3625
3633
.
https://doi.org/10.1021/acs.est.7b05847
.
Klamerth
N.
Malato
S.
Agüera
A.
Fernández-Alba
A.
2013
Photo-Fenton and modified photo-Fenton at neutral pH for the treatment of emerging contaminants in wastewater treatment plant effluents: a comparison
.
Water Research
47
(
2
),
833
840
.
https://doi.org/10.1016/j.watres.2012.11.008
.
Kusic
H.
Peternel
I.
Ukic
S.
Koprivanac
N.
Bolanca
T.
Papic
S.
Bozic
A. L.
2011
Modeling of iron activated persulfate oxidation treating reactive azo dye in water matrix
.
Chemical Engineering Journal
172
(
1
),
109
121
.
https://doi.org/10.1016/j.cej.2011.05.076
.
Leng
Y.
Guo
W.
Shi
X.
Li
Y.
Xing
L.
2013
Polyhydroquinone-coated Fe3O4 nanocatalyst for degradation of rhodamine B based on sulfate radicals
.
Industrial & Engineering Chemistry Research
52
(
38
),
13607
13612
.
https://doi.org/10.1021/ie4015777
.
Leng
Y.
Guo
W.
Shi
X.
Li
Y.
Wang
A.
Hao
F.
Xing
L.
2014
Degradation of Rhodamine B by persulfate activated with Fe3O4: effect of polyhydroquinone serving as an electron shuttle
.
Chemical Engineering Journal
240
,
338
343
.
https://doi.org/10.1016/j.cej.2013.11.090
.
Li
H.
Wan
J.
Ma
Y.
Huang
M.
Wang
Y.
Chen
Y.
2014
New insights into the role of zero-valent iron surface oxidation layers in persulfate oxidation of dibutyl phthalate solutions
.
Chemical Engineering Journal
250
,
137
147
.
https://doi.org/10.1016/j.cej.2014.03.092
.
Li
H.
Wan
J.
Ma
Y.
Wang
Y.
Guan
Z.
2015
Role of inorganic ions and dissolved natural organic matters on persulfate oxidation of acid orange 7 with zero-valent iron
.
RSC Advances
5
(
121
),
99935
99943
.
https://doi.org/10.1039/C5RA16094D
.
Li
J.
Ji
Q.
Lai
B.
Yuan
D.
2017a
Degradation of p-nitrophenol by Fe0/H2O2/persulfate system: optimization, performance and mechanisms
.
Journal of the Taiwan Institute of Chemical Engineers
80
,
686
694
.
https://doi.org/10.1016/j.jtice.2017.09.002
.
Li
X.
Zhou
M.
Pan
Y.
Xu
L.
2017c
Pre-magnetized Fe0/persulfate for notably enhanced degradation and dechlorination of 2, 4-dichlorophenol
.
Chemical Engineering Journal
307
,
1092
1104
.
https://doi.org/10.1016/j.cej.2016.08.140
.
Liang
C.
Guo
Y. Y.
2010
Mass transfer and chemical oxidation of naphthalene particles with zerovalent iron activated persulfate
.
Environmental Science & Technology
44
(
21
),
8203
8208
.
https://doi.org/10.1021/es903411a
.
Lin
C. C.
Chen
Y. H.
2018
Feasibility of using nanoscale zero-valent iron and persulfate to degrade sulfamethazine in aqueous solutions
.
Separation and Purification Technology
194
,
388
395
.
https://doi.org/10.1016/j.seppur.2017.10.073
.
Liu
C.
Wu
B.
2018
Sulfate radical-based oxidation for sludge treatment: a review
.
Chemical Engineering Journal
335
,
865
875
.
https://doi.org/10.1016/j.cej.2017.10.162
.
Liu
C. S.
Shih
K.
Sun
C. X.
Wang
F.
2012
Oxidative degradation of propachlor by ferrous and copper ion activated persulfate
.
Science of the Total Environment
416
,
507
512
.
https://doi.org/10.1016/j.scitotenv.2011.12.004
.
Liu
H.
Bruton
T. A.
Doyle
F. M.
Sedlak
D. L.
2014
In situ chemical oxidation of contaminated groundwater by persulfate: decomposition by Fe (III)-and Mn (IV)-containing oxides and aquifer materials
.
Environmental Science & Technology
48
(
17
),
10330
10336
.
https://doi.org/10.1021/es502056d
.
Liu
C. M.
Diao
Z. H.
Huo
W. Y.
Kong
L. J.
Du
J. J.
2018
Simultaneous removal of Cu2+ and bisphenol A by a novel biochar-supported zero valent iron from aqueous solution: synthesis, reactivity and mechanism
.
Environmental Pollution
239
,
698
705
.
https://doi.org/10.1016/j.envpol.2018.04.084
.
Long
A.
Lei
Y.
Zhang
H.
2014
Degradation of toluene by a selective ferrous ion activated persulfate oxidation process
.
Industrial & Engineering Chemistry Research
53
(
3
),
1033
1039
.
https://doi.org/10.1021/ie402633n
.
Luo
H.
Lin
Q.
Zhang
X.
Huang
Z.
Liu
S.
Jiang
J.
Xiao
R.
Liao
X.
2019
New insights into the formation and transformation of active species in nZVI/BC activated persulfate in alkaline solutions
.
Chemical Engineering Journal
359
,
1215
1223
.
https://doi.org/10.1016/j.cej.2018.11.056
.
Matzek
L. W.
Carter
K. E.
2016
Activated persulfate for organic chemical degradation: a review
.
Chemosphere
151
,
178
188
.
https://doi.org/10.1016/j.chemosphere.2016.02.055
.
Mueller
N. C.
Braun
J.
Bruns
J.
Černík
M.
Rissing
P.
Rickerby
D.
Nowack
B.
2012
Application of nanoscale zero valent iron (NZVI) for groundwater remediation in Europe
.
Environmental Science and Pollution Research
19
(
2
),
550
558
.
https://doi.org/10.1007/s11356-011-0576-3
.
Naim
S.
Ghauch
A.
2016
Ranitidine abatement in chemically activated persulfate systems: assessment of industrial iron waste for sustainable applications
.
Chemical Engineering Journal
288
,
276
288
.
https://doi.org/10.1016/j.cej.2015.11.101
.
Neyens
E.
Baeyens
J.
2003
A review of classic Fenton's peroxidation as an advanced oxidation technique
.
Journal of Hazardous Materials
98
(
1–3
),
33
50
.
https://doi.org/10.1016/S0304-3894(02)00282-0
.
Nidheesh
P. V.
Gandhimathi
R.
2012
Trends in electro-Fenton process for water and wastewater treatment: an overview
.
Desalination
299
,
1
15
.
https://doi.org/10.1016/j.desal.2012.05.011
.
Nie
M.
Yan
C.
Li
M.
Wang
X.
Bi
W.
Dong
W.
2015
Degradation of chloramphenicol by persulfate activated by Fe2+ and zerovalent iron
.
Chemical Engineering Journal
279
,
507
515
.
https://doi.org/10.1016/j.cej.2015.05.055
.
Nie
M.
Yan
C.
Xiong
X.
Wen
X.
Yang
X.
Dong
W.
2018
Degradation of chloramphenicol using a combination system of simulated solar light, Fe2+ and persulfate
.
Chemical Engineering Journal
348
,
455
463
.
https://doi.org/10.1016/j.cej.2018.04.124
.
Niu
C. G.
Wang
Y.
Zhang
X. G.
Zeng
G. M.
Huang
D. W.
Ruan
M.
Li
X. W.
2012
Decolorization of an azo dye Orange G in microbial fuel cells using Fe (II)-EDTA catalyzed persulfate
.
Bioresource Technology
126
,
101
106
.
https://doi.org/10.1016/j.biortech.2012.09.001
.
Nowack
B.
2002
Environmental chemistry of aminopolycarboxylate chelating agents
.
Environmental Science & Technology
36
(
19
),
4009
4016
.
https://doi.org/10.1021/es025683s
.
Oh
W. D.
Lim
T. T.
2019
Design and application of heterogeneous catalysts as peroxydisulfate activator for organics removal: an overview
.
Chemical Engineering Journal
358
,
110
133
.
https://doi.org/10.1016/j.cej.2018.09.203
.
Oh
S. Y.
Kim
H. W.
Park
J. M.
Park
H. S.
Yoon
C.
2009
Oxidation of polyvinyl alcohol by persulfate activated with heat, Fe2+, and zero-valent iron
.
Journal of Hazardous Materials
168
(
1
),
346
351
.
https://doi.org/10.1016/j.jhazmat.2009.02.065
.
Oh
S. Y.
Kang
S. G.
Chiu
P. C.
2010
Degradation of 2, 4-dinitrotoluene by persulfate activated with zero-valent iron
.
Science of the Total Environment
408
(
16
),
3464
3468
.
https://doi.org/10.1016/j.scitotenv.2010.04.032
.
Oh
W. D.
Dong
Z.
Lim
T. T.
2016
Generation of sulfate radical through heterogeneous catalysis for organic contaminants removal: current development, challenges and prospects
.
Applied Catalysis B: Environmental
194
,
169
201
.
https://doi.org/10.1016/j.apcatb.2016.04.003
.
Olmez-Hanci
T.
Arslan-Alaton
I.
2013
Comparison of sulfate and hydroxyl radical based advanced oxidation of phenol
.
Chemical Engineering Journal
224
,
10
16
.
https://doi.org/10.1016/j.cej.2012.11.007
.
Olmez-Hanci
T.
Arslan-Alaton
I.
Genc
B.
2013
Bisphenol A treatment by the hot persulfate process: oxidation products and acute toxicity
.
Journal of Hazardous Materials
263
,
283
290
.
https://doi.org/10.1016/j.jhazmat.2013.01.032
.
Pan
Y.
Zhou
M.
Li
X.
Xu
L.
Tang
Z.
Sheng
X.
Li
B.
2017
Highly efficient persulfate oxidation process activated with pre-magnetization Fe0
.
Chemical Engineering Journal
318
,
50
56
.
https://doi.org/10.1016/j.cej.2016.05.001
.
Park
J. H.
Wang
J. J.
Tafti
N.
Delaune
R. D.
2019
Removal of Eriochrome Black T by sulfate radical generated from Fe-impregnated biochar/persulfate in Fenton-like reaction
.
Journal of Industrial and Engineering Chemistry
71
,
201
209
.
https://doi.org/10.1016/j.jiec.2018.11.026
.
Pouran
S. R.
Aziz
A. A.
Daud
W. M. A. W.
2015
Review on the main advances in photo-Fenton oxidation system for recalcitrant wastewaters
.
Journal of Industrial and Engineering Chemistry
21
,
53
69
.
https://doi.org/10.1016/j.jiec.2014.05.005
.
Pu
M.
Ma
Y.
Wan
J.
Wang
Y.
Huang
M.
Chen
Y.
2014
Fe/S doped granular activated carbon as a highly active heterogeneous persulfate catalyst toward the degradation of Orange G and diethyl phthalate
.
Journal of Colloid and Interface Science
418
,
330
337
.
https://doi.org/10.1016/j.jcis.2013.12.034
.
Pulicharla
R.
Drouinaud
R.
Brar
S. K.
Drogui
P.
Proulx
F.
Verma
M.
Surampalli
R. Y.
2018
Activation of persulfate by homogeneous and heterogeneous iron catalyst to degrade chlortetracycline in aqueous solution
.
Chemosphere
207
,
543
551
.
https://doi.org/10.1016/j.chemosphere.2018.05.134
.
Rao
Y. F.
Qu
L.
Yang
H.
Chu
W.
2014
Degradation of carbamazepine by Fe (II)-activated persulfate process
.
Journal of Hazardous Materials
268
,
23
32
.
https://doi.org/10.1016/j.jhazmat.2014.01.010
.
Ravikumar
K. V. G.
Santhosh
S.
Sudakaran
S. V.
Nancharaiah
Y. V.
Mrudula
P.
Chandrasekaran
N.
Mukherjee
A.
2018
Biogenic nano zero valent iron (Bio-nZVI) anaerobic granules for textile dye removal
.
Journal of Environmental Chemical Engineering
6
(
2
),
1683
1689
.
https://doi.org/10.1016/j.jece.2018.02.023
.
Rodriguez
S.
Vasquez
L.
Costa
D.
Romero
A.
Santos
A.
2014
Oxidation of Orange G by persulfate activated by Fe (II), Fe (III) and zero valent iron (ZVI)
.
Chemosphere
101
,
86
92
.
https://doi.org/10.1016/j.chemosphere.2013.12.037
.
Satapanajaru
T.
Yoo-iam
M.
Bongprom
P.
Pengthamkeerati
P.
2015
Decolorization of Reactive Black 5 by persulfate oxidation activated by ferrous ion and its optimization
.
Desalination and Water Treatment
56
(
1
),
121
135
.
https://doi.org/10.1080/19443994.2014.932710
.
Shang
W.
Dong
Z.
Li
M.
Song
X.
Zhang
M.
Jiang
C.
Feiyun
S.
2019
Degradation of diatrizoate in water by Fe (II)-activated persulfate oxidation
.
Chemical Engineering Journal
361
,
1333
1344
.
https://doi.org/10.1016/j.cej.2018.12.139
.
Song
Q.
Feng
Y.
Liu
G.
Lv
W.
2019
Degradation of the flame retardant triphenyl phosphate by ferrous ion-activated hydrogen peroxide and persulfate: kinetics, pathways, and mechanisms
.
Chemical Engineering Journal
361
,
929
936
.
https://doi.org/10.1016/j.cej.2018.12.140
.
Soubh
A. M.
Baghdadi
M.
Abdoli
M. A.
Aminzadeh
B.
2018
Zero-valent iron nanofibers (ZVINFs) immobilized on the surface of reduced ultra-large graphene oxide (rULGO) as a persulfate activator for treatment of landfill leachate
.
Journal of Environmental Chemical Engineering
6
(
5
),
6568
6579
.
https://doi.org/10.1016/j.jece.2018.10.011
.
Stefaniuk
M.
Oleszczuk
P.
Ok
Y. S.
2016
Review on nano zerovalent iron (nZVI): from synthesis to environmental applications
.
Chemical Engineering Journal
287
,
618
632
.
https://doi.org/10.1016/j.cej.2015.11.046
.
Sun
C.
Zhou
R.
Jianan
E.
Sun
J.
Su
Y.
Ren
H.
2016
Ascorbic acid-coated Fe3O4 nanoparticles as a novel heterogeneous catalyst of persulfate for improving the degradation of 2,4-dichlorophenol
.
RSC Advances
6
(
13
),
10633
10640
.
https://doi.org/10.1039/C5RA22491H
.
Tan
X. F.
Liu
Y. G.
Gu
Y. L.
Xu
Y.
Zeng
G. M.
Hu
X. J.
Liu
S. B.
Wang
X.
Liu
X. M.
Li
J.
2016
Biochar-based nano-composites for the decontamination of wastewater: a review
.
Bioresource Technology
212
,
318
333
.
https://doi.org/10.1016/j.biortech.2016.04.093
.
Teel
A. L.
Ahmad
M.
Watts
R. J.
2011
Persulfate activation by naturally occurring trace minerals
.
Journal of Hazardous Materials
196
,
153
159
.
https://doi.org/10.1016/j.jhazmat.2011.09.011
.
Temiz
K.
Olmez-Hanci
T.
Arslan-Alaton
I.
2016
Zero-valent iron-activated persulfate oxidation of a commercial alkyl phenol polyethoxylate
.
Environmental Technology
37
(
14
),
1757
1767
.
https://doi.org/10.1080/09593330.2015.1131751
.
Tsitonaki
A.
Petri
B.
Crimi
M.
Mosbæk
H.
Siegrist
R. L.
Bjerg
P. L.
2010
In situ chemical oxidation of contaminated soil and groundwater using persulfate: a review
.
Critical Reviews in Environmental Science and Technology
40
(
1
),
55
91
.
https://doi.org/10.1080/10643380802039303
.
Wacławek
S.
Lutze
H. V.
Grübel
K.
Padil
V. V.
Černík
M.
Dionysiou
D. D.
2017
Chemistry of persulfates in water and wastewater treatment: a review
.
Chemical Engineering Journal
330
,
44
62
.
https://doi.org/10.1016/j.cej.2017.07.132
.
Wang
S.
Wang
J.
2017
Comparative study on sulfamethoxazole degradation by Fenton and Fe (II)-activated persulfate process
.
RSC Advances
7
(
77
),
48670
48677
.
https://doi.org/10.1039/C7RA09325 J
.
Wang
S.
Wang
J.
2018
Trimethoprim degradation by Fenton and Fe (II)-activated persulfate processes
.
Chemosphere
191
,
97
105
.
https://doi.org/10.1016/j.chemosphere.2017.10.040
.
Wang
X.
Wang
L.
Li
J.
Qiu
J.
Cai
C.
Zhang
H.
2014
Degradation of Acid Orange 7 by persulfate activated with zero valent iron in the presence of ultrasonic irradiation
.
Separation and Purification Technology
122
,
41
46
.
https://doi.org/10.1016/j.seppur.2013.10.037
.
Wang
C.
Wan
J.
Ma
Y.
Wang
Y.
2016a
Insights into the synergy of zero-valent iron and copper oxide in persulfate oxidation of Orange G solutions
.
Research on Chemical Intermediates
42
(
2
),
481
497
.
https://doi.org/10.1007/s11164-015-2035-0
.
Wang
Q.
Shao
Y.
Gao
N.
Chu
W.
Deng
J.
Shen
X.
Lu
X.
Zhu
Y.
Wei
X.
2016b
Degradation of alachlor with zero-valent iron activating persulfate oxidation
.
Journal of the Taiwan Institute of Chemical Engineers
63
,
379
385
.
https://doi.org/10.1016/j.jtice.2016.03.038
.
Wang
Y.
Chen
S. Y.
Yang
X.
Huang
X. F.
Yang
Y. H.
He
E. K.
Wang
X. Q.
Qiu
R. L.
2017
Degradation of 2, 2′, 4, 4′-tetrabromodiphenyl ether (BDE-47) by a nano zerovalent iron-activated persulfate process: the effect of metal ions
.
Chemical Engineering Journal
317
,
613
622
.
https://doi.org/10.1016/j.cej.2017.02.070
.
Wang
X.
Du
Y.
Liu
H.
Ma
J.
2018
Ascorbic acid/Fe0 composites as an effective persulfate activator for improving the degradation of rhodamine B
.
RSC Advances
8
(
23
),
12791
12798
.
https://doi.org/10.1039/C8RA01506F
.
Wang
H.
Guo
W.
Yin
R.
Du
J.
Wu
Q.
Luo
H.
Liu
B.
Sseguya
F.
Ren
N.
2019a
Biochar-induced Fe (III) reduction for persulfate activation in sulfamethoxazole degradation: insight into the electron transfer, radical oxidation and degradation pathways
.
Chemical Engineering Journal
362
,
561
569
.
https://doi.org/10.1016/j.cej.2019.01.053
.
Wang
L.
Wang
Y.
Ma
F.
Tankpa
V.
Bai
S.
Guo
X.
Wang
X.
2019b
Mechanisms and reutilization of modified biochar used for removal of heavy metals from wastewater: a review
.
Science of the Total Environment
.
https://doi.org/10.1016/j.scitotenv.2019.03.011
.
Wang
S.
Wu
J.
Lu
X.
Xu
W.
Gong
Q.
Ding
J.
Dan
B.
Xie
P.
2019c
Removal of acetaminophen in the Fe2+/persulfate system: kinetic model and degradation pathways
.
Chemical Engineering Journal
358
,
1091
1100
.
https://doi.org/10.1016/j.cej.2018.09.145
.
Wang
S.
Zhao
M.
Zhou
M.
Li
Y. C.
Wang
J.
Gao
B.
Sato
S.
Feng
K.
Yin
W.
Igalavithana
A. D.
Oleszczuk
P.
Wang
X.
Ok
X.
2019d
Biochar-supported nZVI (nZVI/BC) for contaminant removal from soil and water: a critical review
.
Journal of Hazardous Materials
373
,
820
834
.
https://doi.org/10.1016/j.jhazmat.2019.03.080
.
Wei
X.
Gao
N.
Li
C.
Deng
Y.
Zhou
S.
Li
L.
2016
Zero-valent iron (ZVI) activation of persulfate (PS) for oxidation of bentazon in water
.
Chemical Engineering Journal
285
,
660
670
.
https://doi.org/10.1016/j.cej.2015.08.120
.
Wei
D.
Li
B.
Huang
H.
Luo
L.
Zhang
J.
Yang
Y.
Guo
J.
Tang
L.
Zeng
G.
Zhou
Y.
2018
Biochar-based functional materials in the purification of agricultural wastewater: fabrication, application and future research needs
.
Chemosphere
197
,
165
180
.
https://doi.org/10.1016/j.chemosphere.2017.12.193
.
Weng
C. H.
Tsai
K. L.
2016
Ultrasound and heat enhanced persulfate oxidation activated with Fe0 aggregate for the decolorization of CI Direct Red 23
.
Ultrasonics Sonochemistry
29
,
11
18
.
https://doi.org/10.1016/j.ultsonch.2015.08.012
.
Weng
C. H.
Tao
H.
2018
Highly efficient persulfate oxidation process activated with Fe0 aggregate for decolorization of reactive azo dye Remazol Golden Yellow
.
Arabian Journal of Chemistry
11
(
8
),
1292
1300
.
https://doi.org/10.1016/j.arabjc.2015.05.012
.
Weng
C. H.
Ding
F.
Lin
Y. T.
Liu
N.
2015
Effective decolorization of polyazo direct dye Sirius Red F3B using persulfate activated with Fe0 aggregate
.
Separation and Purification Technology
147
,
147
155
.
https://doi.org/10.1016/j.seppur.2015.03.062
.
Wu
X.
Gu
X.
Lu
S.
Qiu
Z.
Sui
Q.
Zang
X.
Miao
Z.
Xu
M.
2015
Strong enhancement of trichloroethylene degradation in ferrous ion activated persulfate system by promoting ferric and ferrous ion cycles with hydroxylamine
.
Separation and Purification Technology
147
,
186
193
.
https://doi.org/10.1016/j.seppur.2015.04.031
.
Wu
X.
Gu
X.
Lu
S.
Qiu
Z.
Sui
Q.
Zang
X.
Miao
Z.
Xu
M.
Danish
M.
2016
Accelerated degradation of tetrachloroethylene by Fe (II) activated persulfate process with hydroxylamine for enhancing Fe (II) regeneration
.
Journal of Chemical Technology & Biotechnology
91
(
5
),
1280
1289
.
https://doi.org/10.1002/jctb.4718
.
Wu
S.
He
H.
Li
X.
Yang
C.
Zeng
G.
Wu
B.
He
S.
Lu
L.
2018
Insights into atrazine degradation by persulfate activation using composite of nanoscale zero-valent iron and graphene: performances and mechanisms
.
Chemical Engineering Journal
341
,
126
136
.
https://doi.org/10.1016/j.cej.2018.01.136
.
Wu
J.
Wang
B.
Blaney
L.
Peng
G.
Chen
P.
Cui
Y.
Deng
S.
Wang
Y.
Huang
J.
Yu
G.
2019
Degradation of sulfamethazine by persulfate activated with organo-montmorillonite supported nano-zero valent iron
.
Chemical Engineering Journal
361
,
99
108
.
https://doi.org/10.1016/j.cej.2018.12.024
.
Xiao
R.
Luo
Z.
Wei
Z.
Luo
S.
Spinney
R.
Yang
W.
Dionysiou
D. D.
2018
Activation of peroxymonosulfate/persulfate by nanomaterials for sulfate radical-based advanced oxidation technologies
.
Current Opinion in Chemical Engineering
19
,
51
58
.
https://doi.org/10.1016/j.coche.2017.12.005
.
Xiao
S.
Cheng
M.
Zhong
H.
Liu
Z.
Liu
Y.
Yang
X.
Liang
Q.
2019
Iron-mediated activation of persulfate and peroxymonosulfate in both homogeneous and heterogeneous ways: a review
.
Chemical Engineering Journal
.
https://doi.org/10.1016/j.cej.2019.123265
Xu
X. R.
Li
X. Z.
2010
Degradation of azo dye Orange G in aqueous solutions by persulfate with ferrous ion
.
Separation and Purification Technology
72
(
1
),
105
111
.
https://doi.org/10.1016/j.seppur.2010.01.012
.
Xu
H.
Tian
W.
Zhang
Y.
Tang
J.
Zhao
Z.
Chen
Y.
2018a
Reduced graphene oxide/attapulgite-supported nanoscale zero-valent iron removal of acid red 18 from aqueous solution
.
Water, Air, & Soil Pollution
229
(
12
),
388
.
https://doi.org/10.1007/s11270-018-4033-5
.
Xu
J.
Zhang
X.
Sun
C.
He
H.
Dai
Y.
Yang
S.
Lin
Y.
Zhan
X.
Li
Q.
Zhou
Y.
2018b
Catalytic degradation of diatrizoate by persulfate activation with peanut shell biochar-supported nano zero-valent iron in aqueous solution
.
International Journal of Environmental Research and Public Health
15
(
9
),
1937
.
https://doi.org/10.3390/technologies8010012
.
Yahya
M. S.
Oturan
N.
El Kacemi
K.
El Karbane
M.
Aravindakumar
C. T.
Oturan
M. A.
2014
Oxidative degradation study on antimicrobial agent ciprofloxacin by electro-Fenton process: kinetics and oxidation products
.
Chemosphere
117
,
447
454
.
https://doi.org/10.1016/j.chemosphere.2014.08.016
.
Yan
J.
Lei
M.
Zhu
L.
Anjum
M. N.
Zou
J.
Tang
H.
2011
Degradation of sulfamonomethoxine with Fe3O4 magnetic nanoparticles as heterogeneous activator of persulfate
.
Journal of Hazardous Materials
186
(
2–3
),
1398
1404
.
https://doi:10.1016/j.jhazmat.2010.12.017
.
Yan
J.
Han
L.
Gao
W.
Xue
S.
Chen
M.
2015
Bidochar supported nanoscale zerovalent iron composite used as persulfate activator for removing trichloroethylene
.
Bioresource Technology
175
,
269
274
.
https://doi.org/10.1016/j.biortech.2014.10.103
.
Yang
S.
Che
D.
2017
Degradation of aquatic sulfadiazine by Fe0/persulfate: kinetics, mechanisms, and degradation pathway
.
RSC Advances
7
(
67
),
42233
42241
.
https://doi.org/10.1039/C7RA07920F
.
Yang
B.
Tian
Z.
Zhang
L.
Guo
Y.
Yan
S.
2015
Enhanced heterogeneous Fenton degradation of Methylene Blue by nanoscale zero valent iron (nZVI) assembled on magnetic Fe3O4/reduced graphene oxide
.
Journal of Water Process Engineering
5
,
101
111
.
https://doi.org/10.1016/j.jwpe.2015.01.006
.
Yu
S.
Gu
X.
Lu
S.
Xue
Y.
Zhang
X.
Xu
M.
Qiu
Z.
Sui
Q.
2018
Degradation of phenanthrene in aqueous solution by a persulfate/percarbonate system activated with CA chelated-Fe (II)
.
Chemical Engineering Journal
333
,
122
131
.
https://doi.org/10.1016/j.cej.2017.09.158
.
Yuan
Y.
Tao
H.
Fan
J.
Ma
L.
2015
Degradation of p-chloroaniline by persulfate activated with ferrous sulfide ore particles
.
Chemical Engineering Journal
268
,
38
46
.
https://doi.org/10.1016/j.cej.2014.12.092
.
Zhang
H.
Wang
Z.
Liu
C.
Guo
Y.
Shan
N.
Meng
C.
Sun
L.
2014
Removal of COD from landfill leachate by an electro/Fe2+/peroxydisulfate process
.
Chemical Engineering Journal
250
,
76
82
.
https://doi.org/10.1016/j.cej.2014.03.114
.
Zhang
B. T.
Zhang
Y.
Teng
Y.
Fan
M.
2015a
Sulfate radical and its application in decontamination technologies
.
Critical Reviews in Environmental Science and Technology
45
(
16
),
1756
1800
.
https://doi.org/10.1080/10643389.2014.970681
.
Zhang
M.
Chen
X.
Zhou
H.
Murugananthan
M.
Zhang
Y.
2015b
Degradation of p-nitrophenol by heat and metal ions co-activated persulfate
.
Chemical Engineering Journal
264
,
39
47
.
https://doi.org/10.1016/j.cej.2014.11.060
.
Zhang
Y.
Tran
H. P.
Du
X.
Hussain
I.
Huang
S.
Zhou
S.
Wen
W.
2017
Efficient pyrite activating persulfate process for degradation of p-chloroaniline in aqueous systems: a mechanistic study
.
Chemical Engineering Journal
308
,
1112
1119
.
https://doi.org/10.1016/j.cej.2016.09.104
.
Zhang
D.
Li
Y.
Tong
S.
Jiang
X.
Wang
L.
Sun
X.
Li
J.
Liu
X.
Shen
J.
2018a
Biochar supported sulfide-modified nanoscale zero-valent iron for the reduction of nitrobenzene
.
RSC Advances
8
(
39
),
22161
22168
.
https://doi.org/10.1039/C8RA04314 K
.
Zhang
K.
Zhou
X.
Zhang
T.
Yu
L.
Qian
Z.
Liao
W.
Li
C.
2018b
Degradation of the earthy and musty odorant 2, 4, 6-tricholoroanisole by persulfate activated with iron of different valences
.
Environmental Science and Pollution Research
25
(
4
),
3435
3445
.
https://doi.org/10.1007/s11356-017-0452-x
.
Zhang
Y.
Zhang
Q.
Dong
Z.
Hong
J.
2018c
Degradation of acetaminophen with ferrous/copperoxide activate persulfate: synergism of iron and copper
.
Water Research
146
,
232
243
.
https://doi.org/10.1016/j.watres.2018.09.028
.
Zhao
J.
Zhang
Y.
Quan
X.
Chen
S.
2010
Enhanced oxidation of 4-chlorophenol using sulfate radicals generated from zero-valent iron and peroxydisulfate at ambient temperature
.
Separation and Purification Technology
71
(
3
),
302
307
.
https://doi.org/10.1016/j.seppur.2009.12.010
.
Zhao
L.
Hou
H.
Fujii
A.
Hosomi
M.
Li
F.
2014
Degradation of 1, 4-dioxane in water with heat-and Fe2+-activated persulfate oxidation
.
Environmental Science and Pollution Research
21
(
12
),
7457
7465
.
https://doi.org/10.1007/s11356-014-2668-3
.
Zhou
L.
Ma
J.
Zhang
H.
Shao
Y.
Li
Y.
2015
Fabrication of magnetic carbon composites from peanut shells and its application as a heterogeneous Fenton catalyst in removal of methylene blue
.
Applied Surface Science
324
,
490
498
.
https://doi.org/10.1016/j.apsusc.2014.10.152
.
Zhou
L.
Zheng
W.
Ji
Y.
Zhang
J.
Zeng
C.
Zheng
Y.
Wang
Q.
Yang
X.
2013
Ferrous-activated persulfate oxidation of arsenic (III) and diuron in aquatic system
.
Journal of Hazardous Materials
263
,
422
430
.
https://doi.org/10.1016/j.jhazmat.2013.09.056
.
Zhu
C.
Fang
G.
Dionysiou
D. D.
Liu
C.
Gao
J.
Qin
W.
Zhou
D.
2016
Efficient transformation of DDTs with persulfate activation by zero-valent iron nanoparticles: a mechanistic study
.
Journal of Hazardous Materials
316
,
232
241
.
https://doi.org/10.1016/j.jhazmat.2016.05.040
.
Zhu
Y.
Yue
M.
Natarajan
V.
Kong
L.
Ma
L.
Zhang
Y.
Zhao
Q.
Zhan
J.
2018
Efficient activation of persulfate by Fe3O4 @ β-cyclodextrin nanocomposite for removal of bisphenol A
.
RSC Advances
8
(
27
),
14879
14887
.
https://doi.org/10.1039/C8RA01696H
.
Zou
X.
Zhou
T.
Mao
J.
Wu
X.
2014
Synergistic degradation of antibiotic sulfadiazine in a heterogeneous ultrasound-enhanced Fe0/persulfate Fenton-like system
.
Chemical Engineering Journal
257
,
36
44
.
https://doi.org/10.1016/j.cej.2014.07.048
.