Fe (II)-activated persulfate oxidation effectively degrades iodoform in water: Influential factors and kinetics analysis

Zijun Dong; Feng Wang; Xiulan Song; Miao Zhang; Chengchun Jiang; Feiyun Sun; Mu Li

doi:10.1016/j.arabjc.2020.01.023

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Original article

13 (

); 5009-5017

doi:

10.1016/j.arabjc.2020.01.023

Fe (II)-activated persulfate oxidation effectively degrades iodoform in water: Influential factors and kinetics analysis

Zijun Dong^{^a,1}, Feng Wang^{^b,1}, Xiulan Song^{^c}, Miao Zhang^{^c}, Chengchun Jiang^{^a}, Feiyun Sun^{^b,⁎}, Mu Li^{^b}

a School of Civil and Environmental Engineering, Shenzhen Polytechnic, Shenzhen 518055, China

b Harbin Institute of Technology (Shenzhen), Shenzhen 518055, China

c College of Environment Science and Technology, Taiyuan University of Technology, Taiyuan 030024, Shanxi, China

⁎Corresponding author. sun_fy@hit.edu.cn (Feiyun Sun)

Received: 2019-10-9, Accepted: 2020-1-28, Published: 2020-04-01

Disclaimer:
This article was originally published by Elsevier and was migrated to Scientific Scholar after the change of Publisher.

1 These authors contributed equally to this work.

Abstract

Abstract

Fe(II) activated persulfate oxidation effectively degrade CHI₃.
Operation condition parameters for CHI₃ degradation were optimized.
The mechanisms of CHI₃ removal degradation by Fe²⁺/PS was unveiled.

Abstract

Iodoform (CHI₃) is one of the disinfection by-products (DBPs) that is formed in the pre-oxidation and disinfection processes of drinking water treatment. In this study, Fe(II)-activated persulfate oxidation (Fe²⁺/PS) was employed to degrade iodoform, the effects of initial reactants concentration, reaction parameters, kinetics model were investigated, and the underlying mechanisms of CHI₃ degradation in Fe²⁺/PS oxidation process was unveiled. The results showed that the mole ratio Fe²⁺/PS of 1:5, initial PS concentration of 15 μmmol/L, and pH of 3.0 were identified as the optimum operating parameters. In addition, a relatively higher temperature could enhance CHI₃ removal and deiodination. The kinetic model has two different reaction steps: a fast one during the first ten minutes of reaction; then followed by a much slower one. Suppression of the reaction by TBA and MeOH showed that the combined effects of SO₄⁻· and ·OH contributed to the degradation of CHI₃, but ·OH played a dominant role. The degradation pathways and the products of total liberated iodine species demonstrate that the applicability of the Fe²⁺/PS oxidation process for CHI₃ degradation. These results indicated that the Fe²⁺/PS oxidation process is an effective advanced oxidation process for CHI₃ removal in water treatment.

Keywords

Fe2+/PS oxidation process

Iodoform degradation

Optimal operating conditions

Kinetics

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1 Introduction

As one of the unregulated iodinated trihalomethanes (I-THMs), iodoform (CHI₃) has been identified as a disinfection by-product (DBP), and which formed through the reactions between natural organic matter (NOM) and reactive iodine species, i.e. hypoiodous acid (HOI) in drinking water treatment process (Bichsel et al., 2000). The highest concentration of I-THMs was found in drinking water treated with chloramines. Hansson et al. reported that a change from chlorination to chloramination lead to the formation of 5 μg/L CHI₃ (I⁻ concentration of 50 μg/L) (Hansson et al., 1987). Krasner et al. observed that the maximum concentration of unregulated I-THMs around 15 μg/L in drinking water, which is relatively lower than the US EPA limitation (< 80 μg/L) (Krasner et al., 2006). According to literature reports, I-THMs are more cytotoxic and genotoxic than their regulated chlorinated and brominated analogs. Richardson et. al. demonstrated that the cytotoxicity of CHI₃ was 146 times and 60 times higher than that of CHCl₃ and CHBr₃, respectively (Richardson et. al., 2008). In addition, CHI₃ could produce bad taste and odor problems in drinking water (Hansson et al., 1987). Hansson and Cancho reported that CHI₃ has a lower odor threshold concentration (0.03–1 μg/L) compared with CHCl₃ and CHBr₃ (300 and 100 μg/L, respectively) (Hansson and Cancho et al., 2001). Therefore, the effective removal of CHI₃ from drinking water is highly expected.

However, there are only a few studies investigated the degradation of CHI₃ in drinking water and wastewater treatment processes, and some researchers recognized that CHI₃ could not be effectively degraded by conventional treatment techniques, such as coagulation, sedimentation, filtration, and ozonation (Amy et al., 1999). Various advanced oxidation processes (AOPs), such as Fenton reactions (Li et al., 2012), UV/H₂O₂ oxidation (Duan et al., 2016), Fe²⁺/H₂O₂ oxidation (Velo-Gala et al., 2014), and Ferrate oxidation (Wang et al., 2018) have been identified as effective ways to remove DBPs and trace organics, as well as to eliminate I-THMs from water. Wang et al. have investigated UV/chlorine advanced oxidation for CHI₃ degradation, and the results showed that over 90% CHI₃ could be effectively removed (Wang et al., 2017a, 2018b). A rapid oxidation of iodide and hypoiodous acid with ferrate in the ferrate/I/HA system was developed by Wang et al., and the experiments displayed that ferrate would rapidly oxidize HOI formed in the I⁻ oxidation process, and when humic acids (HA) existed in the solution, no formation of iodoform and monoiodoacetic acid (MIAA) was observed in the oxidation of iodide with ferrate (from 10 mM to 80 mM) (Wang et al., 2018).

Among the commonly used AOPs oxidants, persulfate (PS) has superior redox potential to effectively degrade organic pollutants (Neta et al., 1988). By comparison of the effect of different AOPs, PS can be activated to produce SO₄⁻· that possessing a similar or even higher oxidation potential (2.5–3.1 V vs. NHE) and a longer half-life period than ·OH (30–40 μs vs 20 ns), SO₄⁻· can transfer long distances to target contaminants and oxidize them more thoroughly. Additionally, those reactions can take place effectively in aqueous systems on a wider pH range (2–8) (Fan et al., 2018, Jin et al., 2018), PS can be activated by ultraviolet (UV) or heat (Gao et al., 2016), and transition metals, such as Cu²⁺, Mn²⁺, Fe²⁺, Co²⁺ and Ag⁺ (Xu et al., 2008), while Fe²⁺ is deemed a less toxic and more affordable agent (Ike et al., 2018). Accordingly, the Fe²⁺/PS reaction system with relatively low energy (14.8 kcal·mol⁻¹) has been widely applied in disinfection (Wordofa et al., 2017, Ike et al., 2018), dewatering activated sludge (Zhen et al., 2012), and some refractory toxic pollutants (Xu et al., 2010). The efficiency of Fe²⁺/PS oxidation characteristics is summarized in Table 1.

Table 1 Brief summary of the efficient of Fe²⁺/PS oxidation characteristics for a series of the target pollutant.

Pollutant	Reaction conditions						Degradation	Reference
Pollutant	pH	Time (min)	Pollutant concentration	Temperature(°C)	Oxidant concentration	Activation agent dose	Degradation	Reference
Acetaminophen	3.0	30	0.05 mM	20	0.8 mM	1.0 mM	81.4%	Wang et al. (2019)
Sulfadiazine	n. a.^a	120	100 μM	25	4.0 mM	1.0 mM	About 100%	Yang et al. (2018)
Trimethoprim	3.0	240	1 mM	25	4.0 mM	4.0 mM	73.4%	Wang et al. (2018)
Chlortetracycline	3.0–4.0	120	1 mM	20	500 mM	1000 mM	76%	Pulicharla et al. (2018)
Sulfamethoxazole	3.0	240	0.05 mM	25	4.0 mM	4.0 mM	100%	Wang et al. (2017)
Atrazine	n. a.	10	20 μM	n. a.	0.4 mM	0.4 mM	About 50%	Bu et al. (2016)
Carbamazepine	3.0	40	0.025 mM	n. a.	1.0 mM	0.125 mM	78%	Rao et al. (2014)
Diuron	4.0–5.0	180	0.09 mM	50	2 mM	0.72 mM	100%	Romero et al. (2010)
Orange G	3.5	60	0.1 mM	20	4.0 mM	4.0 mM	99%	Xu et al. (2010)
Bisphenol A	4.0–5.0	240	0.05 mM	25	2.5 mM	1.25 mM	73%	Zhang et al. (2014)
Domestic wastewater (carbamazepine)	3.0	60	0.5 μM	n. a.	2.5 mM	4.0 mM	100%	Ahmed and Chiron (2014)
Arsenic	n. a.	240	6.6 μM	25	20 μM	20 μM	100%	Zhou et al. (2013)

a n.a.—information not available.

However, to our knowledge, studies on CHI₃ degradation by the Fe²⁺/PS oxidation process are rarely reported. In the present study, Fe²⁺/PS oxidation was employed to remove CHI₃, and the influence of operational parameters on the CHI₃ treatment efficiency was examined to determine the optimum conditions. CHI₃ degradation kinetic model was established to depict the reaction mechanisms and further revealed the CHI₃ degradation pathway. The obtained results will be invaluable for CHI₃ degradation and removal from water and wastewater.

2 Materials and methods

2.1

2.1 Chemicals

All chemicals used in this study were analytical grade. Iodoform (CHI₃, 99%) was purchased from Sigma-Aldrich (USA), and sodium persulfate (Na₂S₂O₈, 99%) was purchased from Aladdin (China). Ferrous sulfate heptahydrate (FeSO₄⋅7H₂O, ≥99%) and sodium thiosulfate (Na₂S₂O₃, ≥99%) were purchased from Xilong Chemical Co., Ltd. Sodium carbonate anhydrous (Na₂CO₃) was from Tianjin Mao Tai Chemical Reagent Factory.

2.2

2.2 Experimental set-up

All experiments were conducted in 250 mL (effective volume: 200 mL) glass reactors. Firstly, a CHI₃ concentration of 5 mg/L was added to the glass reactors, and then added appropriate volumes of ferrous sulfate solution and pure water. The ferrous sulfate solution was freshly prepared, to avoid being oxidized. The initial solution pH was adjusted by using 0.1 M H₂SO₄ and NaOH. A certain volume of sodium persulfate was added, and the reaction was conducted in electric heat thermostat shocks sink at an appropriate temperature. Aliquots of 9 mL were withdrawn at selected time intervals (0 min, 5 min, 10 min, 30 min, and 60 min) and quenched immediately with 1 mL Na₂S₂O₃ after filtered by a filterable membrane. Similarly, the experiments were conducted at the same condition with varied concentrations of SO₄^−•,^•OH, I⁻, IO⁻, and IO₃⁻.

2.3

2.3 Analytical procedures

Agilent 7697 Headspace Sampler and Agilent 6890 gas chromatography (GC) equipped with an electron capture detector (ECD) and DB-1MS fused silica capillary column (30 m × 0.25 mm) were used to determinate iodoform. The temperature program used to separate the analyses was as following procedure: initially column temperature was held at 60 °C for 1 min, increased at 10 °C/min to 150 °C, maintained for 2 min, and then increased at 20 °C/min to 250 °C and held for 5 min. Both of the injector and detector temperature was 260 °C. The temperature of the heating cabinet, sample loop, and transmission line were 60 °C, 90 °C, 110 °C, respectively. Degradation products (I⁻, IO₃⁻) were detected by ion chromatography (IC, Thermo Dionex ICS-1500), which was equipped with a conductivity detector, Ionpac AG9-HC guard column (4 × 50 mm) and Ionpac AS9-HC analytical column (4 × 250 mm). The eluent was 9 mmol/L Na₂CO₃ and the flow velocity was 1 mL/min. The column and detector temperature both were 35 °C. The detection limit was 0.3 μg/L. Samples were filtered by SPE C18 to remove the coexisted organic matter and then injected into ion chromatography. IO⁻ was measured as 4-iodine-2,6-chlorophenesic acid by HPLC equipped with Symmetry C18 column (4.6 mm × 250 mm; 5 μm) and the conditions were as follows: flow rate 1 mL/min, mobile phase acetonitrile/water (1‰ acetic acid) 70%: 30%, detection wavelength 212 mm, sample size 10 μL, and retention 10 min.

The exposure values of sulfate radicals were calculated by the indicator, atrazine (ATZ). The exposure values of hydroxyl radicals could be calculated by the indicator, nitrobenzene (NB). Therefore, free radical exposure can be calculated according to Eqs. (1) and (2) (Xian et al., 2017).

(1)

{(\int [\cdot O H] d t)}_{NB} = - l n (\frac{NB}{{NB}_{0}}) / k_{\cdot O H, N B}

$(\cdot O H, N B)$ ——Secondary reaction rate constants of hydroxyl radical and nitrobenzene (M⁻¹∙s⁻¹).

${(\int [\cdot O H] d t)}_{NB}$ ——The hydroxyl radical exposure value obtained with nitrobenzene as the indicator (M∙s).

(2)

{(\int [S O_{4}^{- \cdot}] d t)}_{ATZ} = \frac{- \ln (\frac{[A T Z]}{AT Z_{0}}) - k_{(\cdot O H, A T Z)} \int [\cdot O H] d t}{k_{S O_{4}^{-} \cdot, A T Z}}

$k_{S O_{4}^{- \cdot}, A T Z}$ ——Secondary reaction constants of sulfate-free radical and atrazine.

$k_{(\cdot O H, A T Z)}$ ——Secondary reaction constants of hydroxyl radical and atrazine.

${(\int [S O_{4}^{- \cdot}] d t)}_{ATZ}$ ——The sulfate-free radical exposure value obtained with atrazine as a tracer (M∙s).

2.4

2.4 Kinetic model

The kinetic model of the CHI₃ degradation in the Fe²⁺/PS system can be expressed as shown in Eq. (3) (Frontistis et al., 2011).

(3)

\frac{t}{1 - (C / C_{0})} = ρ + δ t \underset{}{⟺} \frac{d / (C / C_{0})}{dt} = \frac{- ρ}{{(ρ + δ t)}^{2}}

where C and C₀ are the substrate concentration at time t and 0,

ρ

and

δ

are two characteristic kinetic constants. The reciprocal of

ρ

corresponds to the initial reaction rate (i.e. at t = 0), while the reciprocal of

ρ

corresponds to the maximum conversion that can be achieved at the end of the reaction.

3 Results and discussion

3.1

3.1 Effects of reactants concentration

3.1.1

3.1.1 PS concentration

The effects of initial PS concentration and Fe²⁺/PS mole ratio on CHI₃ degradation were studied, as shown in Fig. 1. The effects of PS concentration on CHI₃ degradation was investigated at a pH of 3.0, a temperature of 25 °C and a constant Fe²⁺/PS mole ratio of 1:5. Fig. 1(a) showed the concentration of CHI₃ variation under the different initial Na₂S₂O₈ concentrations, 2, 4, 6, 8, 10, and 15 μmol/L, all of the reactions activated by sufficient Fe²⁺. It can be clearly observed that an increased PS concentration from 2 to 15 μmol/L leads to a continuous decrease of CHI₃ concentration. The CHI₃ degradation rate increased rapidly within the first 10 min and then went smoothly. A higher initial PS concentration induced a greater removal efficiency of CHI₃, and its removal efficiency reached 83% when the PS concentration was 15 μmol/L, during which IO₃⁻ could be detected. IO₃⁻ is one stable sink of iodine species in the drinking water treatment process because it is proven to be non-toxic and it can be converted to I⁻ by glutathione quickly (Zhen et al., 2016). As shown in Fig. 1(b), the amount of IO₃⁻ generated has a positive correlation with an initial concentration of PS. When the initial dosage of PS reach to 15 μmol/L, the amount of IO₃⁻ could reach to 0.5 μmol/L with the decreasing of CHI₃ concentration in the bulk solution, indicating there was the occurrence of deiodination reaction.

The effects of initial PS concentration on the CHI3 degradation and ROS production.

CHI₃ was believed to be degraded by the oxidative free radicals SO₄⁻⋅ and ⋅OH generated in the Fe²⁺/PS oxidation system. Fig. 1(c, d) showed the exposure values of SO₄⁻⋅ and ⋅OH during oxidation, and it was obvious that SO₄⁻⋅ generated in the Fe²⁺/PS system was increased with the increasing of PS concentration, and the highest SO₄⁻⋅ reach to about 2.4 μmol/L. However, as the concentration of PS raised, the number of hydroxyl groups insignificantly changed. Due to the Fe²⁺/PS oxidation mainly attributed to the combined effects of SO₄⁻⋅ and ⋅OH (Shang et al., 2019), the persulfate oxidation was more beneficial and cost-effective in enhancing CHI₃ degradation, owing to the lower requirement for pH adjustment (Zhen et al., 2012). Generally, SO₄⁻⋅ was thought to be the primary ROS during the oxidation of PS, and an increasing PS concentration produced more SO₄⁻⋅, thus improved the oxidation capacity of the reaction system and enhanced CHI₃ removal efficiency.

3.1.2

3.1.2 Fe²⁺/PS mole ratio

CHI₃ degradation efficiency was also affected by the Fe²⁺/PS mole ratio. Under the condition of a temperature of 25 °C, pH of 3.0, and an initial PS dosage of 4 μmol/L, the effect of Fe²⁺/PS mole ratio on CHI₃ degradation was studied by changing Fe²⁺ dosage from 0.1 to 5.0 mmol/L, which corresponded to n(Fe²⁺)/n(PS) of 1:20, 1:10, 1:5, 1:2, 1:1 and 2:1. The variations of CHI₃ degradation efficiency and the IO₃⁻ concentration with the reaction were shown in Fig. 2. It can be observed from Fig. 2(a) that CHI₃ removal efficiency increased rapidly with increasing mole ratio of Fe²⁺/PS from 1:20 to 1:5, and then reached to the maximum once the n(Fe²⁺)/n(PS) increased to 1:5. After the continuous increase of n(Fe²⁺)/n(PS) to 2:1, CHI₃ removal efficiency began to decrease. As shown in Fig. 2(b), a similar profile of IO₃⁻ generated with the change of n(Fe²⁺)/n(PS) could be found. As the generated IO₃⁻ increased firstly with the increasing of mole ratio from 1:20 to 1:5, and then began to decrease with the increasing of mole ratio from 1:5 to 2:1. This result displayed that either an insufficient or an excessive Fe²⁺ in the bulk solution was unfavorable for CHI₃ degradation and deiodination when the PS dosage was fixed.

The effects of Fe2+/PS mole ratio on the CHI3 degradation and ROS production.

The effective exposure values of SO₄⁻⋅ and ⋅OH in the reaction showed in Fig. 2(c, d). When the Fe²⁺/PS ratio increased from 1:20 to 1:5, the produced SO₄⁻⋅ and ⋅OH were promoted significantly. Afterward, a further increase of n(Fe²⁺)/n(PS) above 1:5 resulted in a decreasing trend in SO₄⁻⋅ and ⋅OH concentration. It was observed that the change of SO₄⁻⋅ was coincident with CHI₃ degradation efficiency under varied n(Fe²⁺)/n(PS). The Fe²⁺ dosage in an appropriate range around 1:5 could increase SO₄⁻⋅ generated from Fe²⁺/PS system, which could eventually enhance its oxidizing capability and CHI₃ degradation efficiency. However, once the Fe²⁺ dosage increased to above 1:5, the excessive Fe²⁺ would react with SO₄⁻• (Eq. (4)), while the SO₄⁻• in a high concentration could also self-quenched via Eq. (5) (Rodriguez et al., 2014) to induce insufficient SO₄⁻• in the bulk. This negative consumption of SO₄⁻• limited the oxidation ability of the reaction system, leading to a low CHI₃ degradation efficiency.

(4)

{Fe}^{2 +} + {SO}_{4}^{-} ∙ \to {Fe}^{3 +} + {SO}_{4}^{2 -} K = 4.6 \times 10^{9} m o l L^{- 1}

(5)

{SO}_{4}^{-} ∙ + {SO}_{4}^{-} ∙ \to S_{2} O_{8}^{2 -} K = 8.9 \times 10^{8} m o l L^{- 1}

3.2

3.2 Effects of operation condition parameters on CHI₃ degradation

3.2.1

3.2.1 Initial solution pH

The effects of initial pH on CHI₃ degradation were investigated under a pH of 3, 5, 7, 9 and 11, with a PS dosage of 2 μmol/L, n(Fe²⁺)/n(PS) of 1:5 and an ambient temperature of 25 °C. As presented in Fig. 3(a), when the initial pH was kept at 3.0, more than 52% CHI₃ could be degraded. However, with the increasing of initial pH, the CHI₃ degradation efficiency began to decrease continuously, to even as low as 20% when the initial pH reached to 11. This negative correlation between CHI₃ degradation efficiency and initial pH ranged from 3 to 11 may be attributed to an increased pH resulted in an obviously decreased IO₃⁻ generation (Fig. 3 (b)), leading to a rather low deiodination capability.

The effects of the initial pH on the CHI3 degradation and ROS production.

Fig. 3 (c, d) demonstrated that an acid condition was positive to SO₄⁻• production, as more SO₄⁻• could be generated with the decreasing of pH from 7 to 3. In contrast, under an alkaline condition, less SO₄⁻• was generated, while more ^•OH might be produced that could slightly increase the solution pH. The reasons may be that in alkaline conditions, persulfate was easily decomposed to generate SO₄⁻•, which generated ^•OH by reacting with OH^– in solution (Rodriguez et al., 2014), as shown in Eq. (6). Since the oxidation-reduction potential of ^•OH was lower than SO₄⁻•, the oxidizing ability of the system would be dropped off, and decreased CHI₃ removal efficiency. In addition, SO₄²⁻ existed in the reaction system could also consume ·OH by reacting with it and weaken the oxidizing ability of the system.

(6)

{SO}_{4}^{-} ∙ + {OH}^{-} \to ∙ O H + {SO}_{4}^{2 -}

3.2.2

3.2.2 Temperature

The effects of the temperature of 25, 30, 35, 40 and 45 °C on CHI₃ removal rate were studied with a PS dosage of 2 μmol/L, n(Fe²⁺)/n(PS) of 1:5 and a pH of 3.0. The CHI₃ removal rate and IO₃⁻ variation was shown in Fig. 4. It was found that CHI₃ degradation efficiency increased remarkably with the temperature increased from 25 °C to 45 °C at the beginning of 10 min reaction. More than 70% of CHI₃ could be removed at a temperature above 45 °C. Similarly, IO₃⁻ generation increased with elevated temperature (Fig. 2(b)) in the Fe²⁺/PS process. The reaction was an endothermic reaction between Fe²⁺ and PS (Shang et al., 2019), therefore, a high temperature can enhance PS reaction to produce more SO₄⁻• in this process, and lead to a remarkable degradation efficiency and a high degree of deiodination directly.

The effects of operating temperature on the CHI3 degradation and ROS production.

Fig. 4(c, d) showed more SO₄⁻• could be produced at a high temperature, but there was no obvious impact on ^•OH. This was due to the fact that thermal could also activate PS to produce SO₄⁻•, and more SO₄⁻• were generated with the increased temperature (Han et al., 2015). In addition, owing to that the reaction was an endothermic reaction, an increased temperature could also accelerate the reaction that leads to an enhanced degradation affection and high deiodination rate (Frontistis, 2018). The coupling of thermal treatment with the Fe²⁺/PS process enhanced the oxidation ability and improved CHI₃ removal efficiency.

3.3

3.3 Degradation kinetics and paths analysis

CHI₃ degradation in the present work likely has two different reaction steps: a fast degradation rate during the first 10 min reaction, followed by a much slower rate. Fitting of the data is excellent, the respective kinetic constants and the coefficients of regression (r²) were presented in Table 2. Increased n(Fe²⁺)/n(PS) mole ratio would increased the initial rate, and the maximum conversion rate was achieved at the n(Fe²⁺)/n(PS) mole ratio of 1:5. The removal rate of CHI₃ also affected by acidity and alkalinity of the solution. Under an acid condition, CHI₃ has a higher removal efficiency, which was similar to the Fenton process that depends on the ions forms in solution. As the temperature increased, the activity of the Fe²⁺/PS oxidation process gradually elevated, and the relative concentration of PS, iron and initial CHI₃ concentration influenced the degradation efficiency. As a matter of fact, the PS to CHI₃ concentration ratio at a constant catalyst concentration is a critical parameter to keep that the reaction would not be limited by the number of hydroxyl radicals generated (i.e. at very low ratios) or it was not hindered by auto-scavenging reactions (i.e. at excessive ratios) (Frontistis et al., 2011; Chan and Chu, 2003).

Table 2 Kinetic modeling of CHI₃ degradation by Fe²⁺/PS oxidation.

Run	n (Fe²⁺)/n (PS)	pH	T(°C)	ρ⁻¹ (min⁻¹)	δ⁻¹ (%)	r² (%)
1	1:20	3.0	25	0.09	37	98.6
2	1:10			0.15	47	98.3
3	1:5			0.15	52	98.7
4	1:2			0.11	37	98.6
5	1:1			0.08	33	99.8
6	2:1			0.06	14	99.5
7	1:5	3.0	25	0.16	52	98.6
8		5.0		0.15	47	99.5
9		7.0		0.15	37	99.6
10		9.0		0.06	25	99.6
11		11		0.05	16	99.3
12	1:5	3.0	25	0.15	52	98.6
13			30	0.16	53	99.0
14			35	0.17	55	97.6
15			40	0.26	59	96.1
16			45	0.27	68	93.8

A scavenging experiment was conducted to examine the role of radicals in CHI₃ degradation by using methanol that has significant inhibitory effects on both SO₄⁻• and ^•OH, and by using TBA that has a strong inhibitory effect on •OH and a weak constraint on SO₄⁻ (Wordofa et al., 2017). As shown in Fig. 5(a), more than 53% CHI₃ could be degraded without any radical scavenger. When there was a coexistence of methanol or TBA with a ratio of 500:1 and 2000:1 (methanol/TBA), the CHI₃ degradation efficiency decreased to 23% and 48%, 20% and 33%, respectively. It was evidenced that CHI₃ degradation in the Fe²⁺/PS oxidation process was mainly attributed to the combined function of SO₄⁻• and ^•OH. Generally, SO₄⁻• was thought to be the primary ROS during the oxidation of PS. Nevertheless, the contribution of SO₄⁻• to CHI₃ degradation could not be overestimated, as there were a series of reactions listed as Eqs. (7)–(17) (Wang et al., 2017a, 2017b).

(7)

{Fe}^{2 +} + S_{2} O_{8}^{2 -} \to {SO}_{4}^{-} ∙ + F e (I I I) + {SO}_{4}^{2 -} K < 30 M^{- 1} s^{- 1}

(8)

S O_{4}^{-} ∙ + H_{2} O \to ∙ O H + {SO}_{4}^{2 -} + H^{+} K < 3 \times 10^{3} M^{- 1} s^{- 1}

(9)

S O_{4}^{-} ∙ + O H^{-} \to ∙ O H + {SO}_{4}^{2 -} K = (6.5 \pm 1.0) \times 10^{7} M^{- 1} s^{- 1}

(10)

S_{2} O_{8}^{2 -} + ∙ O H \to S O_{4}^{-} ∙ + H S O_{4}^{-} + 1 / 2 O_{2} K = 1.2 \times 10^{7} M^{- 1} s^{- 1}

(11)

S O_{4}^{-} ∙ + ∙ O H \to H S O_{4}^{-} + 1 / 2 O_{2} K = 1.0 \times 10^{10} M^{- 1} s^{- 1}

(12)

S_{2} O_{8}^{2 -} + S O_{4}^{-} ∙ \to S_{2} O_{8}^{2 -} + {SO}_{4}^{2 -} K = 6.1 \times 10^{5} M^{- 1} s^{- 1}

(13)

S O_{4}^{-} ∙ + S O_{4}^{-} ∙ \to S_{2} O_{8}^{2 -} K = 4.0 \times 10^{8} M^{- 1} s^{- 1}

(14)

S_{2} O_{8}^{2 -} + 2 H_{2} O \to 2 {SO}_{4}^{2 -} + H O_{2}^{-} + 3 H^{+}

(15)

2 S_{2} O_{8}^{2 -} + 2 H_{2} O \to 3 {SO}_{4}^{2 -} + S O_{4}^{-} ∙ + O_{2}^{-} ∙ + 4 H^{+}

(16)

2 S_{2} O_{8}^{2 -} + H O_{2}^{-} \to {SO}_{4}^{2 -} + S O_{4}^{-} ∙ + O_{2}^{-} ∙ + H^{+}

(17)

S_{2} O_{8}^{2 -} + O_{2}^{-} ∙ \to {SO}_{4}^{2 -} + S O_{4}^{-} ∙ + O_{2}

CHI3 degradation in Fe2+/PS oxidation process: (a) The effects of different inhibitors; (b) the degradation products examination.

The degradation products of CHI₃ by Fe²⁺/PS process were analyzed, and the concentration changes of iodine species (I, IO₃⁻ and HOI) with reaction time were illustrated. As shown in Fig. 5(b), the concentration of CHI₃ kept a decreasing trend from the beginning to 30 min. While, the concentration of I⁻ increased rapidly within the first 30 min and then reached a plateau between 30 and 60 min, and it contributed 52.2% liberated iodine species. Within the first 20 min, HOI and I₂ concentration increased to 3.8 μM and then kept relatively steady gradually to 4.2 μM at 30 min. At the end of the reaction, IO₃⁻ was detected and took a small proportion (< 1%) of total liberated iodine species. Taurog et al. studied the conversion of iodate in vivo, which proved that iodine species could quickly convert to I⁻ by glutathione. Thus, iodine species were common stable sink in the drinking water treatment process, because it has nontoxic. However, compared with the heat-activated persulfate oxidation process, I⁻ is presumed to transform to IO₃⁻ ultimately in the presence of excessive SO₄⁻•. Meanwhile, in a UV process for CHI₃ degradation, there was no IO₃⁻ observed (Xiao et al., 2014), and less of 2% total iodine species was found in terms of IO₃⁻ during UV/H₂O₂ process (Xiao et al., 2015). Compared with heat-activated persulfate oxidation, UV, UV/H₂O₂, and UV/chlorination processes, Fe²⁺/PS oxidation is a promising AOP technology for the CHI₃ removal from water.

4 Conclusion

Fe²⁺/PS oxidation process is examined to be an efficient treatment technique for CHI₃ degradation, and over 83% CHI₃ was removed in 60 min. The mole ratio Fe²⁺/PS of 1:5, PS concentration of 15 μmmol/L, and an initial pH of 3.0 was verified to be the optimum operating condition parameters. The reaction temperature displayed a positive correlation with CHI₃ removal and deiodination efficiency when the temperatures ranging from 25 °C to 45 °C. Suppression of the reaction by TBA and MeOH showed that the combined effects of SO₄⁻· and ·OH contributed to the degradation of CHI₃, while ·OH played a dominant role. The kinetic model, pathways and the products of total liberated iodine species demonstrate the applicability of the Fe²⁺/PS oxidation system for CHI₃ degradation and provides a new insight for DBPs degradation and removal in an aqueous environment.

Acknowledgment

This research was supported by the Grant Project Supported by Guangdong Natural Science Foundation No. 2017A030313285 from the Shenzhen Science and Technology Funding Project [grant number JCYJ20160406162038258], Program of International S&T Cooperation [grant number 2016YFE0123400]; National Natural Science Foundation of China [grant numbers 51678183, 51408149]; Shenzhen Science and Technology Project [grant numbers JSGG2017041401900541]. The authors appreciate the thoughtful and constructive comments from the anonymous reviewers.

Declaration of Competing Interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

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Introduction

Materials and methods

Chemicals
Experimental set-up
Analytical procedures
Kinetic model

Results and discussion

Effects of reactants concentration
Effects of operation condition parameters on CHI3 degradation
Degradation kinetics and paths analysis

Conclusion

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[1] Ahmed M.M., Chiron S., . Solar photo-Fenton like using persulphate for carbamazepine removal from domestic wastewater. Water Res.. 2014;48:229-236.
[Google Scholar]

[2] Amy G.L., Siddiqui M.S., . Strategies to control bromate and bromide. American water works association; 1999.

[3] Bichsel Y., Von Gunten U., . Formation of iodo-trihalomethanes during disinfection and oxidation of iodide-containing waters. Environ. Sci. Technol.. 2000;34:2784-2791.
[Google Scholar]

[4] Bu L.J., Shi Z., Zhou S.Q., . Modeling of Fe(II)-activated persulfate oxidation using atrazine as a target contaminant. Sep. Purif. Technol.. 2016;169:59-65.
[Google Scholar]

[5] Cancho B., Fabrellas C., Diaz A., Ventura F., Galceran M.T., . Determination of the odor threshold concentrations of iodinated trihalomethanes in drinking water. J. Agric. Food. Chem.. 2001;49:1881-1884.
[Google Scholar]

[6] Chan K.H., Chu W., . Modeling the reaction kinetics of Fenton's process on the removal of atrazine. Chemosphere. 2003;51(4):305-311.
[Google Scholar]

[7] Duan X., He X., Wang D., Mezyk S.P., Otto S.C., Marfil-Vega R., Mills M.A., . Decomposition of iodinated pharmaceuticals by UV-254 nm-assisted advanced oxidation processes. J. Hazard. Mater.. 2016;323:489-499.
[Google Scholar]

[8] Fan J., Gu L., Wu D., Liu Z., . Mackinawite (FeS) activation of persulfate for the degradation of p-chloroaniline: surface reaction mechanism and sulfur-mediated cycling of iron species. Chem. Eng. J.. 2018;333:657-664.
[Google Scholar]

[9] Frontistis Z., Xekoukoulotakis N.P., Hapeshi E., Venieri D., Fatta-Kassinos D., Mantzavinos D., . Fast degradation of estrogen hormones in environmental matrices by photo-fenton oxidation under simulated solar radiation. Chem. Eng. J.. 2011;178:175-182.
[Google Scholar]

[10] Frontistis Z., . Degradation of the nonsteroidal anti-inflammatory drug piroxicam by iron activated persulfate: the role of water matrix and ultrasound synergy. Int. J. Environ. Res. Public Health. 2018;15(11):2600.
[Google Scholar]

[11] Gao Y., Zhang Z., Li S., Liu J., Yao L., Li Y.X., . Insights into the mechanism of heterogeneous activation of persulfate with a clay/iron-based catalyst under visible LED light irradiation. Appl. Catal. B. 2016;185:22-30.
[Google Scholar]

[12] Han D., Wan J., Ma Y., Wang Y., Li Y., Li D., . New insights into the role of organic chelating agents in Fe(II) activated persulfate processes. Chem. Eng. J.. 2015;269:425-433.
[Google Scholar]

[13] Hansson R.C., Henderson M.J., Jack P., Taylor R.D., . Iodoform taste complaints in chloramination. Water Res.. 1987;21:1265-1271.
[Google Scholar]

[14] Ike I.A., Linden K., Orbell J.D., Duke M., . Critical review of the science and sustainability of persulphate advanced oxidation processes. Chem. Eng. J.. 2018;338:651-669.
[Google Scholar]

[15] Jin Q.Q., Zhang S., Wen T., Wang J., Gu P.C., Zhao G.X., . Simultaneous adsorption and oxidative degradation of Bisphenol A by zero-valent iron/iron carbide nanoparticles encapsulated in N-doped carbon matrix. Environ. Pollut.. 2018;243:218-227.
[Google Scholar]

[16] Krasner S.W., Weinberg H.S., Richardson S.D., Pastor S.J., Chinn R., Sclimenti M.J., . Occurrence of a new generation of disinfection byproducts. Environ. Sci. Technol.. 2006;40:7175-7185.
[Google Scholar]

[17] Li W., Nanaboina V., Zhou Q., Korshin G.V., . Effects of Fenton treatment on the properties of effluent organic matter and their relationships with the degradation of pharmaceuticals and personal care products. Water Res.. 2012;46:403-412.
[Google Scholar]

[18] Neta P., Huie R.E., Ross A.B., . Rate constants for reactions of inorganic radicals in aqueous solution. J. Phys. Chem. Ref. Data. 1988;17:1027-1284.
[Google Scholar]

[19] Pulicharla R., Drouinaud R., Brar S.K., Drogui P., Proulx F., Verma M., . Activation of persulfate by homogeneous and heterogeneous iron catalyst to degrade chlortetracycline in aqueous solution. Chemosphere. 2018;207:543-551.
[Google Scholar]

[20] Rao Y.F., Qu L., Yang H.S., Chu W., . Degradation of carbamazepine by Fe(II)-activated persulfate process. J. Hazard. Mater.. 2014;268:23-32.
[Google Scholar]

[21] Richardson S.D., Fasano F., Ellington J.J., Crumley F.G., Buettner K.M., Evans J.J., . Occurrence and mammalian cell toxicity of iodinated disinfection byproducts in drinking water. Environ. Sci. Technol.. 2008;42:8330-8338.
[Google Scholar]

[22] Rodriguez S., Vasquez L., Costa D., Romero A., Santos A., . Oxidation of Orange G by persulfate activated by Fe(II), Fe(III) and zero valent iron (ZVI) Chemosphere. 2014;101:86-92.
[Google Scholar]

[23] Romero A., Santos A., Vicente F., Concepción G., . Diuron abatement using activated persulphate: effect of pH, Fe(II) and oxidant dosage. Chem. Eng. J.. 2010;162:257-265.
[Google Scholar]

[24] Shang W., Dong Z., Li M., Song X., Zhang M., Jiang C., . Degradation of diatrizoate in water by Fe(II)-activated persulfate oxidation. Chem. Eng. J.. 2019;361:1333-1344.
[Google Scholar]

[25] Velo-Gala Inmaculada, López-Peñalver Jesús J., Sánchez-Polo Manuel, Rivera-Utrilla José, . Comparative study of oxidative degradation of sodium diatrizoate in aqueous solution by H₂O₂/Fe²⁺, H₂O₂/Fe³⁺, Fe (VI) and UV, H₂O₂/UV, K₂S₂O₈/UV. Chem. Eng. J.. 2014;241:504-512.
[CrossRef] [Google Scholar]

[26] Wang A.Q., Lin Y.L., Xu B., Hu C.Y., Xia S.J., Zhang T.Y., . Kinetics and modeling of iodoform degradation during uv/chlorine advanced oxidation process. Chem. Eng. J.. 2017;323:312-329.
[Google Scholar]

[27] Wang L., Kong D., Ji Y., Lu J., Yin X., Zhou Q., . Transformation of iodide and formation of iodinated by-products in heat activated persulfate oxidation process. Chemosphere. 2017;181:400-408.
[Google Scholar]

[28] Wang S.L., Wu J.F., Lu X.Q., Xu W.X., Gong Q., Ding J.Q., . Removal of acetaminophen in the Fe²⁺/persulfate system: kinetic model and degradation pathways. Chem. Eng. J.. 2019;358:1091-1100.
[Google Scholar]

[29] Wang X.S., Liu Y.L., Huang Z.S., Wang L., Wang Y.C., Li Y.T., . Rapid oxidation of iodide and hypoiodous acid with ferrate and no formation of iodoform and monoiodoacetic acid in the ferrate/I^-/HA system. Water Res.. 2018;144:592-602.
[Google Scholar]

[30] Wordofa D.N., Walker S.L., Liu H., . Sulfate radical-induced disinfection of pathogenic escherichia coli O157:H7 via iron-activated persulfate. Environ. Sci. Technol. Lett.. 2017;4:154-160.
[Google Scholar]

[31] Xian L., Shao Y., Gao N., Chen J.X., Peng F.Y., . Investigation of clofibric acid removal by UV/persulfate and UV/chlorine processes: Kinetics and formation of disinfection byproducts during subsequent chlor(am)ination. Chem. Eng. J.. 2017;331:364-371.
[Google Scholar]

[32] Xiao Y., Fan R., Zhang L., Yue J., Webster R.D., Lim T.T., . Photodegradation of iodinated trihalomethanes in aqueous solution by UV 254 irradiation. Water Res.. 2014;49:275-285.
[Google Scholar]

[33] Xiao Y., Zhang L., Yue J., Webster R.D., Lim T.T., . Kinetic modeling and energy efficiency of UV/H₂O₂ treatment of iodinated trihalomethanes. Water Res.. 2015;75:259-269.
[Google Scholar]

[34] Xu X.R., Li X.Z., . Degradation of azo dye Orange G in aqueous solutions by persulfate with ferrous ion. Sep. Purif. Technol.. 2010;72:105-111.
[Google Scholar]

[35] Xu X., Ye Q., Tang T., Wang D., . Hg⁰ oxidative absorption by K₂S₂O₈ solution catalyzed by Ag⁺ and Cu²⁺. J. Hazard. Mater.. 2008;158:410-416.
[Google Scholar]

[36] Yang H., Zhuang S., Hu Q., Hu L.T., Yang P., Au C.T., . Competitive reactions of hydroxyl and sulfate radicals with sulfonamides in Fe²⁺/S₂O₈^2- system: reaction kinetics, degradation mechanism and acute toxicity. Chem. Eng. J.. 2018;339:32-41.
[Google Scholar]

[37] Zhang Y.Q., Xie X.F., Huang S.B., Liang H.Y., . Effect of chelating agent on oxidation rate of aniline in ferrous ion activated persulfate system at neutral pH. J. Cent. South Univ.. 2014;21:1441-1447.
[Google Scholar]

[38] Zhen G., Lu X., Zhao Y., Chai X., Niu D., . Enhanced dewater ability of sewage sludge in the presence of Fe(II)-activated persulfate oxidation. Bioresour. Technol.. 2012;116:259-265.
[Google Scholar]

[39] Zhen W., Lin Y.L., Xu B., Xia S.J., Zhang T.Y., Gao N.Y., . Degradation of iohexol by UV/chlorine process and formation of iodinated trihalomethanes during post-chlorination. Chem. Eng. J.. 2016;283:1090-1096.
[Google Scholar]

[40] Zhou L., Zheng W., Jia Y., Zhang J., Zeng C., Zhang Y., Wang Q., Yang X., . Ferrousactivated persulfate oxidation of arsenic(III) and diuron in aquatic system. J. Hazard. Mater.. 2013;263:422-430.
[Google Scholar]

Fe (II)-activated persulfate oxidation effectively degrades iodoform in water: Influential factors and kinetics analysis

Abstract

Abstract