Environ Sci Pollut Res DOI 10.1007/s11356-014-3718-6

RESEARCH ARTICLE

Degradation and dechlorination of pentachlorophenol by microwave-activated persulfate Chengdu Qi & Xitao Liu & Wei Zhao & Chunye Lin & Jun Ma & Wenxiao Shi & Qu Sun & Hao Xiao

Received: 30 June 2014 / Accepted: 10 October 2014 # Springer-Verlag Berlin Heidelberg 2014

Abstract The degradation performance of pentachlorophenol (PCP) by the microwave-activated persulfate (MW/PS) process was investigated in this study. The results indicated that degradation efficiency of PCP in the MW/PS process followed pseudo-first-order kinetics, and compared with conventional heating, microwave heating has a special effect of increasing the reaction rate and reducing the process time. A higher persulfate concentration and reaction temperature accelerated the PCP degradation rate. Meanwhile, increasing the pH value and ionic strength of the phosphate buffer slowed down the degradation rate. The addition of ethanol and tert-butyl alcohol as hydroxyl radical and sulfate radical scavengers proved that the sulfate radicals were the dominant active species in the MW/PS process. Gas chromatography-mass spectrometry (GC-MS) was employed to identify the intermediate products, and then a plausible degradation pathway involving dechlorination, hydrolysis, and mineralization was proposed. The acute toxicity of PCP, as tested with Photobacterium phosphoreum, Vibrio fischeri, and Vibrio qinghaiensis, was negated quickly during the MW/PS process, which was in agreement with the nearly complete mineralization of PCP. Responsible editor: Angeles Blanco Electronic supplementary material The online version of this article (doi:10.1007/s11356-014-3718-6) contains supplementary material, which is available to authorized users. C. Qi : X. Liu (*) : C. Lin : J. Ma : W. Shi : Q. Sun State Key Laboratory of Water Environment Simulation, School of Environment, Beijing Normal University, Beijing 100875, China e-mail: [email protected] W. Zhao Institute of Scientific and Technical Information of China, Beijing 100038, China H. Xiao College of Chemistry, Beijing Normal University, Beijing 100875, China

These results showed that the MW/PS process could achieve a high mineralization level in a short time, which provided an efficient way for PCP elimination from wastewater. Keywords Pentachlorophenol . Persulfate . Microwave . Dechlorination . Toxicity

Introduction Pentachlorophenol (PCP) was introduced by humans in the 1930s as a preservative for wood and is still used extensively in agricultural and industrial applications as a fungicide, bactericide, herbicide, molluscicide, algaecide, and insecticide (Crosby 1981). The wide use of PCP has exposed the aquatic and terrestrial environment to pollution (Zheng et al. 2011). Owing to its acute and chronic toxicity and carcinogenic nature, PCP has been listed as a “priority pollutant” by the US-EPA (Hayward 1998) and the WHO (WHO 2011). Therefore, it is important to develop reliable and effective methods to eliminate PCP from aqueous solutions. Current methods proposed for the removal of PCP mainly include physisorption (Deng et al. 2009), biodegradation (Garg et al. 2013; Pu and Cutright 2007), chemical reductive dechlorination (Jia et al. 2012; Kim and Carraway 2000), and advanced oxidation processes (AOPs) (Liu et al. 2004; Rodgers et al. 1999; Sung et al. 2012; ThanhThuy et al. 2013; Zimbron and Reardon 2009). However, physisorption only transfers pollutants from the aqueous phase to the surface of adsorbents. Biodegradation usually requires a long residence time for microorganisms to degrade the pollutant because they are affected by PCP toxicity. Chemical reductive dechlorination has been successfully utilized to treat PCP, e.g., by zero-valent iron (Kim and Carraway 2000) or Pd0/Fe0 (Jia et al. 2012). AOPs, which involve the generation of nonselective hydroxyl radicals (·OH), are effective for eliminating

Environ Sci Pollut Res

refractory organic pollutants from aqueous solutions. The common methods of AOPs used to treat PCP-contaminated wastewater include photo-catalysis (ThanhThuy et al. 2013), microwave degradation (Liu et al. 2004), ozonation (Sung et al. 2012), Fenton’s oxidation (Zimbron and Reardon 2009), and electrochemical oxidation (Rodgers et al. 1999). Recently, the activated persulfate (PS) oxidation process appears to be one of the most promising technologies to remove recalcitrant organic pollutants from water due to its environmental friendliness and cost effectiveness. In aqueous solution, the oxidant persulfate anion (S2O82−, E0 =2.01 V) can be activated to generate an even stronger oxidant known as sulfate radicals (SO4−·, E0 =2.60 V) (Ahmad et al. 2013; Dogliotti and Hayon 1967; Furman et al. 2010; House 1962; Kolthoff and Miller 1951). S2 O8 2− þ heat=hυ→2 SO4 − ⋅

employed in effectively decomposing sulfamethoxazole (Qi et al. 2014) and perfluorooctanoic acid (Lee et al. 2009) in water. In this study, microwave heating was applied to activate persulfate for the destruction of PCP. The effects of some important reaction parameters, such as heating mode, reaction temperature, initial persulfate concentration, phosphatebuffered pH value, and phosphate buffer ionic strength, on the degradation performance were investigated. Gas chromatography-mass spectrometry (GC-MS) was applied to determine the intermediates, and the degradation pathway of PCP was proposed accordingly. Additionally, the toxicity of the reaction solution was evaluated using three luminescent bacteria, Photobacterium phosphoreum, Vibrio fischeri, and Vibrio qinghaiensis.

ð1Þ Materials and methods

S2 O8 2− þ Menþ →SO4 − ⋅ þ Meðnþ1Þþ þSO4 2−

2 S2 O8 2− þ 5 OH− →4 SO4 2− þ ⋅ OH þ ⋅ O2 − þ 2 H2 O

ð2Þ

ð3Þ

Among the methods for persulfate activation, thermal activation may be the easiest way to control free radical production (Tsitonaki et al. 2010). Another advantage of thermal activation is that the solubility of organic pollutants in aqueous solution may be increased, which allows more pollutants to be oxidized, as the activated persulfate oxidation occurs in the aqueous phase (Costanza et al. 2010). The technique of thermally activated persulfate oxidation has been applied for the degradation of methyl tertiary butyl ether (MTBE) (Huang et al. 2002), 1,4-dioxane (Zhao et al. 2014), and chlorinated solvents (Liang et al. 2007; Waldemer et al. 2007). Compared with conventional heating, microwave heating can reduce the activation energy, shorten the reaction time, and enhance the reaction rate (Yang et al. 2009). Microwaves induce molecular motion by nonionizing radiation, namely ion migration and dipole rotation, and heat is generated through friction and collision. Therefore, microwave heating possesses the ability to provide the necessary energy for the reaction medium quickly (Costa et al. 2009). In the process of microwave heating, approximately 85 % of the electrical energy can be transformed into thermal energy. It has been observed that the reaction rate of microwave heating-activated persulfate degradation of polyethylene oxide is much faster than that obtained by electric heating activation (Vijayalakshmi et al. 2005). Furthermore, the microwaveactivated persulfate (MW/PS) process has been successfully

Materials All chemical reagents and organic solvents were at least analytical grade and used as received without further purification. Pentachlorophenol (98.5 %), methanol, ethanol, tertbutyl alcohol, and dichloromethane were obtained from J&K Scientific (Beijing, China). The derivatization reagent N,OBis(trimethylsilyl)trifluoroacetamide/trimethylchlorosilane (BSTFA/TMCS) was supplied by AccuStandard (New Haven, CT, USA). Sodium persulfate and other inorganic chemicals used were purchased from Sinopharm (Beijing, China). All solutions were prepared using water with a resistance of 18.2 MΩ cm−1 from a Milli-Q Integral 5 system (EMD Millipore, Billerica, MA, USA). Experimental procedure A stock solution of PCP (37.5 μM, 10 mg L−1) was prepared in Milli-Q deionized water prior to every batch experiment. This solution was not pH-adjusted except for the experiments evaluating the effects of pH (phosphate-buffered pH value 4.4, 7.0, and 9.0 with ionic strength ([H2PO4−] +[HPO42−]= 100 mM)) and phosphate-buffer (PB) ionic strength ([H2PO4−]+[HPO42−]=10, 50, and 100 mM, respectively). The resulting solution was stirred overnight at room temperature until PCP was completely dissolved and was then stored in the dark at 4 °C in amber glass bottles with Teflon-lined caps. Organic cosolvents were never used. Sodium persulfate solutions were prepared before its use. A closed-vessel microwave digestion system (ETHOS-A, Milestone Inc., Shelton, CT, USA) operating at 2.45 GHz with maximum power of 1,200 W and temperature of 300 °C was used for microwave heating (MW). Temperature settings are controlled by sensors automatically. Once the temperature

Environ Sci Pollut Res

reaches the prespecified temperature, the microwave control system will dynamically adjust the microwave power accordingly. Thirty milliliters of PCP solution (37.5 μM) with 0.5 mL of an appropriate concentration persulfate solution was held in a 100 mL tetrafluorometoxil (TFM) vessel with sleeve for the microwave treatment. The temperature program was set to raise the temperature of the solution from 25 to 60, 90, or 130 °C in 3 min and then hold it constant throughout the entire duration of irradiation. At the end, the treated samples were quickly placed in an ice bath (4 °C) to quench the reaction by chilling and set aside for analysis. The time of entire duration of irradiation excluding the 3 min of the temperature-raising period was defined as the microwave irradiation time. Conventional heating (CH) degradation experiments were carried out in a constant temperature water bath shaker unit (THZ-82A, Jiangsu Jingtan Ronghua Instrumental Factory, Jiangsu, China), which can be operated in temperatures ranging from ambient temperature to 100±0.5 °C. Tests were conducted in 150-mL Erlenmeyer flasks with cover containing 100-mL PCP solution (37.5 μM). Prior to the addition of persulfate, the flasks were preheated in the water bath for 1 h to allow the solutions to reach the prespecified reaction temperature. The reaction was initiated once 0.5 mL of sodium persulfate solution of appropriate concentration was added. During the treatment, rapid shaking (∼100 rpm) ensured complete solution mixing. At each sampling time, 1.6 mL of sample was collected and stored in a 1.8-mL sample bottle which was quickly placed in an ice bath (4 °C) to quench the reaction by chilling prior to further analysis. All degradation experiments were conducted with duplicate samples.

set at 1.0 mL min−1. The dechlorination efficiency can be expressed as (moles of Cl− formed)/(moles of chlorine content in initial PCP). The total organic carbon (TOC) content was detected by a TOC analyzer (Liqui TOC, elementar, Germany), which is based on a combustion catalytic oxidation method, using a highly sensitive multichannel nondispersive infrared detector (NDIR). All pH measurements were made with a basic pH meter PB-10 (Sartorius, China). Toxicity assay The acute toxicity was evaluated with a rapid detector for water toxicity (BHP9511, Hamamatsu Photonics, Beijing, China), a commercial bioassay based on the inhibition of bioluminescence emitted by the marine bacteria P. phosphoreum, V. fischeri, and freshwater luminescent bacterium V. qinghaiensis (Hamamatsu Photonics, Beijing, China). The light emission intensity was measured after a sample contact period of 15 min (DIN 2007) and used to calculate the inhibition according to Eq. (4). Inhibition ¼ ðL0 − LÞ=L0

ð4Þ

where L0 is the luminescent intensity of blank solution and L is the luminescent intensity of a sample. To ensure the accuracy, each sample was tested in duplicates. The pH of samples was adjusted to 7 by addition of 1 M NaOH before the test.

Sample analysis Results and discussion The concentration of PCP was measured by a highperformance liquid chromatography (HPLC, Ultimate 3000, Dionex, Sunnyvale, CA, USA) equipped with a Zorbax Eclipse Plus C18 column (150 mm×4.6 mm, 5 μm, Agilent, Santa Clara, CA, USA) and UV detector with the wavelength of 320 nm. The column temperature was set at 25 °C and the injection volume was 20 μL. The mobile phase was a mixture of methanol and 1 % (w/w) acetic acid (85:15, v/v), and flow rate was 1.0 mL min−1. The intermediates during the MW/PS process of PCP degradation were identified using a Varian 4000 GC-MS system according to the method reported by Niu et al. (2013). Concentrations of chloride ion (Cl−) and sulfate ion (SO42 − ) were analyzed with an ion chromatograph system (ICS1100, Dionex, USA) equipped with a guard column and a separation column (Dionex IonPac AS19 Analytical, 4 mm× 250 mm, USA). The injection volume was 20 μL and the column temperature was controlled at 35 °C. The mobile phase was composed of 20 mM KOH, and the flow rate was

Effect of the reaction temperature and heating mode The reaction temperature is considered to be one of the most important parameters to influence the rate of the degradation reaction. It is also a major factor to affect the operating cost. The degradation performance of PCP under different heating modes and temperatures, with or without persulfate, is shown in Fig. 1. A certain degree of degradation efficiency (20.8 %) was achieved by reacting PCP (37.5 μM) without persulfate under microwave heating at 130 °C for 40 min as a blank control. Adding persulfate (112.5 μM) into the PCP solution without microwave heating for 40 min achieved only 0.6 % degradation efficiency (data not shown). This result implies that S2O82− is not an efficient oxidant for degrading PCP at room temperature. Under microwave or conventional heating at 60 and 90 °C, the efficiency of PCP degradation was improved by the increased temperature. The PCP degradation efficiency can be simulated with pseudo-first-order kinetics,

Environ Sci Pollut Res

Fig. 1 Evolution of the PCP degradation (a) in the heat-activated persulfate process, dechlorination (b), and sulfate concentration (c) in the MW/PS process ([PCP]=37.5 μΜ, [PS]=112.5 μΜ, no pH adjustment)

and the rate constants (kobs) for microwave heating were 0.0225 min−1 (R2 =0.99) at 60 °C and 0.0408 min−1 (R2 = 0.99) at 90 °C. Those rate constants for conventional heating were 0.0095 min−1 (R2 =0.99) at 60 °C and 0.0130 min−1 (R2 =0.99) at 90 °C. The values of kobs under microwave heating were approximately 2.3–3.1 times of those obtained under conventional heating at the same temperatures (60 and 90 °C). These results indicated that persulfate could be transformed into SO4−· radicals by microwave heating resulting in a specific effect on the degradation of PCP which was similar to our previous study (Qi et al. 2014). Under microwave heating at 130 °C, 89.8 % of PCP was degraded rapidly. Afterward, the degradation efficiency increased only slightly when a longer reaction time was employed. The corresponding rate constant in the constant temperature stage was 0.0265 min−1 (R2 =0.95) at 130 °C. Higher temperature accelerated the PCP degradation, but an extremely high temperature would level off the reaction rate because a large quantity of radical oxidants were released rapidly to consume most of the remaining persulfate ions (House 1962). SO4 − ⋅ þ S2 O8 2− →SO4 2− þ S2 O8 − ⋅

ð5Þ

Furthermore, the dechlorination efficiencies of PCP at different microwave heating temperatures with persulfate can be

found in Fig. 1b. At 60 °C, the dechlorination efficiency increased at a slow pace as the reaction progressed. At 90 and 130 °C, the dechlorination efficiencies increased at much faster paces and longer reaction times. After 40 min of MW irradiation, the dechlorination efficiencies increased to 11.6, 67.2, and 99.4 % at 60, 90, and 130 °C, respectively. Moreover, the amounts of SO42− formed during degradation of PCP under microwave heating are given in Fig. 1c. At 60 and 90 °C, SO42− accumulated at a slower pace, and its concentration kept increasing with a longer irradiation time. As expected, at 130 °C, almost all S2O82− species were converted into SO42− within 10 min; the SO42− concentration increased slightly with a longer reaction time. After 40 min of MW heating, SO42− concentrations increased to 0.031, 0.213, and 0.225 mM at 60, 90, and 130 °C, respectively.

Effect of the initial persulfate concentration The effect of the initial persulfate concentration on the degradation performance of PCP by MW/PS process was investigated at five persulfate concentrations (0, 37.5, 187.5, 375, and 750 μM) with an initial PCP concentration of 37.5 μM without pH adjustment at 90 °C. In addition to the above experiments, an individual control experiment of 37.5 μM

Environ Sci Pollut Res

PCP with 750 μM PS was conducted under conventional heating at 90 °C. Figure 2a displays the impact of initial persulfate concentration on the degradation efficiency of PCP. As the initial persulfate concentration increased, the degradation rate of PCP also increased correspondingly. At an initial persulfate concentration of 750 μM, PCP was completely removed after 40 min, while 83.9, 65.4, 3.8, and 3.0 % PCP remained after 40 min with initial persulfate concentrations of 0, 37.5, 187.5, and 375 μM, respectively. On the other hand, control experiments with conventional heating of 90 °C showed 5.8 % PCP that remained after 40-min reaction. The rate constants (kobs) are 0.0058 min−1 (R2 =0.98) for an initial persulfate concentration of 0, 0.0097 min −1 (R 2 = 0.97) at 37.5 μM, 0.0523 min−1 (R2 =0.98) at 187.5 μM, 0.0863 min−1 (R2 = 0.99) at 375 μM, and 0.1526 min−1 (R2 =0.99) at 750 μM under microwave heating and 0.0707 min−1 (R2 =0.96) at 750 μM under conventional heating. The results imply that as initial concentrations of persulfate rose, more sulfate radicals were generated, resulting in more rapid degradation of PCP. However, the increment of initial PCP degradation rate was not proportional to the increase of persulfate concentration, possibly due to competition with organic intermediates generated during the degradation of PCP and potential

quenching of sulfate radicals by residual persulfate and sulfate radicals themselves (McElroy and Waygood 1990). SO4 − ⋅ þ SO4 − ⋅ →S2 O8 2−

ð6Þ

Meanwhile, the dechlorination efficiencies and SO42− concentrations during MW/PS process at different initial persulfate concentrations are given in Fig. 2b, c, respectively. As the initial persulfate concentration increased, the dechlorination efficiencies and SO42− concentrations also increased correspondingly. The degradation process of PCP can release Cl− and consume SO4−· radicals which convert into SO42− eventually (Zhao et al. 2010). After 40 min of MW irradiation and the addition of 0, 37.5, 187.5, 375, and 750 μM PS, the dechlorination efficiencies increased from 48.0 to 56.0, 82.0, 90.0, and 96.5 %, when the SO42− concentrations increased from 0 to 0.073, 0.367, 0.737, and 1.418 mM, respectively. Oxidative mechanism in the MW/PS system It has been thoroughly proved that alcohols with and without α-hydrogens have different reactivities and rate constants in their reactions with radical species. Ethanol (EtOH, containing

Fig. 2 Evolution of the PCP degradation (a), dechlorination (b), and sulfate concentration (c) in the MW/PS process at different initial PS concentration ([PCP]=37.5 μΜ, T=90 °C, no pH adjustment)

Environ Sci Pollut Res

an α-hydrogen) reacts with ·OH and SO4−· radicals at high and comparable rates, and the rate constants for the reactions with ·OH and SO4−· radicals are (1.2–2.8)×109 M−1 s−1 and (1.6–7.7)×107 M−1 s−1, respectively. However, tert-butyl alcohol (TBA, without an α-hydrogen) has much different reaction rate constants, and the rate constant for ·OH radicals ((3.8–7.6)×108 M−1 s−1) is 418–1,900 times greater than that for SO4−· radicals ((4–9.1)×105 M−1 s−1) (Liang and Su 2009). Therefore, the addition of the two alcohols into the degradation system will change the degradation rate of PCP, and we can thereby evaluate the reaction mechanisms and the species that play dominant roles in the MW/PS system. As shown in Fig. 3, approximately 94.8 % of PCP was removed in the absence of alcohol additives. However, 64.6 and 34.3 % reductions in the degradation efficiency were observed in the presence of EtOH and TBA, respectively. The rate constants decreased from 0.0408 min−1 (R2 =0.99) in the absence of alcohols to 0.007 min−1 (R2 =0.97) and 0.0222 min−1 (R2 = 0.95) in the presence of EtOH and TBA, respectively. These results showed that the reaction was severely quenched by adding EtOH which acts as an ·OH and SO4−· radicals scavenger, while the addition of TBA only moderately influenced the degradation rate of PCP. This phenomenon in the presence of EtOH provided evidence for the involvement of hydroxyl or sulfate radicals in the mechanism of the MW/PS system. Furthermore, because TBA is also a scavenger of hydroxyl radicals, the degradation of PCP in the presence of TBA should be primarily ascribed to the formation of sulfate radical. Therefore, we believe that sulfate radicals are the dominant active species responsible for the degradation of PCP in the MW/PS system.

1.0

[PCP]/[PCP]0

0.8

0.6

0.4

0.2

0.0

MW/PS MW/PS/TBA MW/PS/EtOH 0

5

10

15

20 35

40

Time (min)

Fig. 3 Evolution of the PCP degradation in the presence of alcohols ([PCP]=37.5 μΜ, [PS]=112.5 μΜ, [Alcohol]=0.1125 M, T=90 °C, no pH adjustment)

Effect of the initial pH and buffer The solution pH plays an important role in the degradation of organic contaminants by the activated persulfate process (Liang et al. 2007; Xu et al. 2014). The degradation performance of PCP (37.5 μM) by PS (112.5 μM) at 90 °C was investigated at three pH values (phosphate buffer at pH 4.4, 7.0, and 9.0, respectively). As it can be observed in Fig. 4a, the most efficient system is the acidic condition (pH 4.4). The rate constants (kobs) clearly showed the superiority of the acidic (0.0378 min −1 (R 2 = 0.97)) pH condition over neutral (0.0265 min−1 (R2 =0.99)) and basic (0.0225 min−1 (R2 = 0.99)) pH conditions. At high pH values, hydroxyl ions in the solution may scavenge SO4−· radicals, which is then transformed into ·OH radicals. However, compared to SO4−· radicals, ·OH radicals have relatively shorter lifetime in aqueous solutions and are therefore unable to effectively degrade PCP because they have less opportunity to get close to the PCP molecules. SO4 − ⋅ þ OH− → ⋅ OH þ SO4 2−

ð7Þ

Consequently, this may slow down PCP degradation in basic pH compared to neutral pH mediums. In addition, acid catalysis initiates a rapid transformation of PS into SO4−· radicals, which, at high concentrations, significantly favor PCP degradation. Under acidic conditions, decomposition of PS into SO4−· radicals can be further acid-catalyzed, and the acid catalysis increases as the pH decreases (Nie et al. 2014). Therefore, at lower pH values, higher efficiency of PCP degradation was observed. The dechlorination efficiencies and SO42− concentrations during the MW/PS process at different pH values are indicated in Fig. 4b, c, respectively. As the initial pH value increased, the dechlorination efficiencies and SO42− concentrations decreased correspondingly. The degradation of PCP can release Cl− and promote the decomposition of PS. After 40 min of MW irradiation, the dechlorination efficiencies decreased from 48.4 to 42.9 and 39.3 %, and the SO42− concentrations decreased from 0.196 to 0.150 and 0.133 mM, as the pH was increased from 4.4 to 7.0 and 9.0, respectively. In contrast to PCP degradation, dechlorination, and SO42− formation, the pH values of the phosphatebuffered solutions were almost unchanged (Fig. 4d). They dropped to 4.3, 7.0, and 8.8 from the initial pH values of 4.4, 7.0, and 9.0, respectively, which could be attributed to the buffering qualities of the phosphate-buffered solution as reported by Liang et al. (2007). To assess the effect of phosphate buffer ionic strengths on the degradation performance of PCP, additional reactions of PCP (37.5 μM) with PS (112.5 μM) at 90 °C were investigated at four ionic strengths ([H2PO4−]+[HPO42−]=0, 10, 50, and 100 mM, respectively). As indicated in Fig. 5a, phosphate

Environ Sci Pollut Res 1.0

50

(a)

(b) 40

Dechlorination (%)

[PCP]/[PCP]0

0.8

0.6

0.4

pH 4.4 pH 7.0 pH 9.0

0.2 0

30

20

10

5

10

15

20 35

40

0 0

5

Time (min)

15

20 35

40

20 35

40

9

(c)

8

(d)

0.150

pH value

SO42- concentration (mM)

0.225

10

Time (min)

7

6

0.075

5 0.000 0

5

10

15

20 35

0

40

5

Time (min)

10

15

Time (min)

Fig. 4 Evolution of the PCP degradation (a), dechlorination (b), sulfate concentration (c), and pH value (d) in the MW/PS process at different initial pH values ([PCP]=37.5 μΜ, [PS]=112.5 μΜ, T=90 °C)

buffer additives have a distinct effect on the degradation rate of PCP. As the initial PB ionic strength ([I]PB) increased, the degradation efficiencies of PCP decreased markedly. The rate constant (kobs) in buffer-free PCP solution (0.0408 min−1 (R2 = 0.99) at [I]PB = 0 mM) is higher than at [I]PB = 10 mM (0.0366 min−1 (R2 =0.98)), [I]PB =50 mM (0.0309 min−1 (R2 =0.99)), and [I]PB =100 mM (0.0265 min−1 (R2 =0.99)). Accordingly, a high buffer capacity has to be avoided. However, a moderate buffer concentration should be maintained for two reasons: to prevent a serious pH drop and to mimic natural effluent conditions, where a significant ionic strength is always expected. It has been demonstrated that some of the sulfate and hydroxyl radicals generated through

SO4 − ⋅ þ HPO4 2− → SO4 2− þ HPO4 − ⋅

ð8Þ

SO4 − ⋅ þ H2 PO4 − → SO4 2− þ H2 PO4 ⋅

ð9Þ

⋅ OH þ HPO4 2− =H2 PO4 − → OH− þ HPO4 − ⋅=H2 PO4 ⋅

1.0

ð10Þ

7.3

(a)

7.2

0.8 7.1 0.6

pH value

[PCP]/[PCP]0

Fig. 5 Evolution of the PCP degradation (a) and pH value (b) at different PB ionic strength ([PCP]=37.5 μΜ, [PS]= 112.5 μΜ, T=90 °C)

thermal PS activation react with phosphate species (Lipczynska-Kochany et al. 1995; Maruthamuthu and Neta 1978).

0.4

[I]PB=0 mM [I]PB=50 mM

4.0

[I]PB=100 mM

0.0 0

5

5.0 4.5

[I]PB=10 mM

0.2

7.0

3.5 10

15

Time (min)

20 35

40

(b) 0

5

10

15

Time (min)

20 35

40

Environ Sci Pollut Res

Fig. 6 Proposed pathway for the oxidation of PCP by MW/PS process

Consequently, with the presence of phosphate buffer, a smaller number of ·OH and SO4−· radicals are available for PCP oxidation, which can influence the PCP degradation rate. The generated phosphate radicals have a lower oxidation capability than the ·OH and SO4−· radicals; therefore, they cannot be considered as secondary oxidants of PCP. The overall pH changes are reported in Fig. 5b. After 40 min of MW irradiation, the pH values of solutions remained nearly constant, except for the buffer-free one with a variation of 1.4 pH units.

propionic acid, butanedioic acid, and 2,3,4,5,6pentahydroxyhexanal can be formed, and these were detected. The C=C bond cleavage might be due to the nucleophilic attack on quinone products by sulfate radicals and H abstraction. These intermediates were further degraded to smaller molecule compounds and finally mineralized to CO2 and H2O.

The degradation pathway of PCP

While our results demonstrate that MW/PS process can effectively degrade PCP, additional experiments were conducted to examine TOC removal and toxicity toward V. fischeri, P. phosphoreum, and V. qinghaiensis. However, owing to the more than 10 % error of TOC analyses for low carbon concentrations (e.g., only 3.9 ppm TOC in the 37.5 μM PCP solution) (Yalfani et al. 2011), these results were not sufficiently reliable to allow a quantitative calculation of the TOC removal during the MW/PS process. Meanwhile, because the EC50 value of PCP is very low for the bioluminescent bacteria (e.g., 2.64 ppm for P. phosphoreum (Kaiser and Ribo 1988)), 1.0

V. Fischeri P. Phosphoreum V. Qinghaiensis TOC

0.8

40

0.6

0.4

0

[TOC]/[TOC]0

80

Inhibition (%)

GC-MS was employed to identify the intermediate products formed during the MW/PS process (see GC-MS spectra in Fig. S1–S6). Based on the results and the previous studies available (Anipsitakis et al. 2005; Peyton 1993; Xu et al. 2013; Zhao et al. 2010), a plausible degradation pathway was proposed in Fig. 6. The sulfate radicals generated from microwave-activated persulfate easily approached the aromatic ring. It attacked PCP followed by elimination to form the carbon-centered radicals via electron transfer from the organic compounds (Anipsitakis et al. 2005), and then the carboncentered radicals led to the formation of hydroxylated products, such as the m-dihydroxybenzene and mtrihydroxybenzene detected in this study, through hydrolysis (Peyton 1993). Chlorine atoms in the benzene ring were also released during this process, which led to the accumulation of Cl− during the MW/PS process as illustrated in previous sections. Because PCP and its dechlorination products, such as dichlorinated and trichlorinated derivatives, are potential precursors of PCDD/Fs (Vallejo et al. 2013), additional analysis was performed to examine the possible formation of PCDD/Fs in the MW/PS process. The results (see in Table. S1) indicated that there were no detectable PCDD/Fs formed in the MW/PS process. Meanwhile, quinone can be produced through hydrogen abstraction via the decomposition of PS from hydroxylated products. Furthermore, following cleavage of quinone’s C=C bond(s) and ring opening,

Toxicity assessments

0.2 -40

0

5

10

15

20 35

40

0.0

Time (min)

Fig. 7 Evolution of the TOC removal and toxicity of PCP solution during MW/PS process ([PCP]=0.375 mΜ, [PS]=7.50 mΜ, T=90 °C)

Environ Sci Pollut Res

the initial and residual PCP content in the solution should be low to observe toxicity changes during the treatment. Taking these factors into account, we chose to increase the initial PCP concentration to 0.375 mM (100 ppm) with 7.5 mM PS. The TOC was measured at the original solution state, and then the solution was diluted 10 times before running the toxicity assays. Figure 7 shows TOC removal and relative inhibition values for three luminescent bacteria during the MW/PS process. The results obtained with V. fischeri, P. phosphoreum, and V. qinghaiensis separately demonstrated that the solutions of 37.5 μM PCP, 0.75 mM PS, and 1.5 mM Na2SO4 alone led to inhibition percentages of 77.8/−21.3/−20.8, 73.7/−22.1/ −31.1, and 99.0/−2.2/−2.2 %, respectively (data not shown). The negative toxicity values indicate that the luminescent intensity of the sample was higher than blank solution, suggesting that certain concentrations of PS or Na2SO4 promote the growth of luminescent bacteria. The relative inhibition of the unheated PCP with PS solution assessed by V. fischeri, P. phosphoreum, and V. qinghaiensis increased slightly to 80.4, 79.8, and 99.2 %, respectively. As mentioned above, no distinct PCP degradation was achieved after 40-min treatment by PS at room temperature. The increased inhibition of the three bacteria should be related to the addition of PS, resulting in synergistic toxicity from PCP and PS. When the solution was heated to 90 °C under microwave irradiation, the relative inhibition values for V. fischeri and P. phosphoreum decreased sharply to 0.4 and 4.0 %, while a 97.9 % inhibition value still held for V. qinghaiensis. The inhibition continued to decrease to −57.4, −41.9, and 50.7 % after 5-min treatment. Afterward, the inhibition values for V. fischeri and P. phosphoreum increased slightly, while the value for V. qinghaiensis continued to decrease. At the end of the 40min treatment, the relative inhibition values for V. fischeri, P. phosphoreum, and V. qinghaiensis were about −31.6, −29.7, and −4.3 %, respectively. This promoting effect suggests that certain nutrients of luminescent bacteria such as small carboncontaining compounds are produced by the degradation of PCP. The inconsistent toxicity results observed between the three different species were mainly due to different natural characteristics of the species tested, such as the intrinsic sensitivity, as well as their ability to metabolize the tested compounds. In addition, the inconsistency may also arise from the different bioavailability of the tested compounds between freshwater and saltwater (Kuang et al. 2013). Consistent with the changes of toxicity values during the MW/PS process, the TOC decreased quickly to nearly 1.7 % of its original content.

Conclusion The effects of heating mode, reaction temperature, persulfate concentration, initial phosphate buffer pH value, and

phosphate buffer ionic strength on degradation performance of PCP in the MW/PS process were investigated. Ethanol and tert-butyl alcohol were used to distinguish the dominant radicals in the MW/PS process. PCP degradation products were identified by GC-MS to elucidate the pathway of PCP degradation, and the acute toxicity of the PCP reaction mixture solution during the MW/PS process was also evaluated. The following conclusions could be drawn from this work: &

& &

&

&

Increasing the reaction temperature (60–130 °C) and persulfate concentration (0–750 μM) resulted in significantly enhanced degradation rates of PCP. Compared with conventional heating, microwave heating displayed a special effect in accelerating the degradation rates. Increasing the phosphate buffer pH value (4.4–9.0) and phosphate buffer ionic strength (0–100 mM) slowed down the degradation rates. The addition of ethanol and tert-butyl alcohol as hydroxyl radicals and sulfate radicals scavengers proved that the sulfate radicals were the dominant radicals in the MW/PS process. The major degradation products were identified by GCMS as m-dihydroxybenzene, m-trihydroxybenzene, propionic acid and butanedioic acid, and the degradation pathway of PCP by the MW/PS process was proposed to include dechlorination, hydrolysis, and mineralization. The acute toxicity tests with V. fischeri, P. phosphoreum, and V. qinghaiensis indicated that 10-times diluted unheated PCP and PCP-persulfate mixture solutions were toxic but became nontoxic after a 40-min MW/PS reaction, which is in agreement with the nearly complete mineralization of PCP.

Acknowledgments The study was supported by the Ministry of Science and Technology (Project No. 2013AA06A305) and the Ministry of Environmental Protection of China (Project No. 201309044).

References Ahmad M, Teel AL, Watts RJ (2013) Mechanism of persulfate activation by phenols. Environ Sci Technol 47:5864–5871 Anipsitakis GP, Dionysiou DD, Gonzalez MA (2005) Cobalt-mediated activation of peroxymonosulfate and sulfate radical attack on phenolic compounds. Implications of chloride ions. Environ Sci Technol 40:1000–1007 Costa C, Santos VHS, Araujo PHH, Sayer C, Santos AF, Fortuny M (2009) Microwave-assisted rapid decomposition of persulfate. Eur Polym J 45:2011–2016 Costanza J, Otaño G, Callaghan J, Pennell KD (2010) PCE oxidation by sodium persulfate in the presence of solids. Environ Sci Technol 44: 9445–9450 Crosby DG (1981) Environmental chemistry of pentachlorophenol. Pure Appl Chem 53:1051–1080

Environ Sci Pollut Res Deng S, Ma R, Yu Q, Huang J, Yu G (2009) Enhanced removal of pentachlorophenol and 2,4-D from aqueous solution by an aminated biosorbent. J Hazard Mater 165:408–414 DIN (2007) Water quality-determination of the inhibitory effect of water samples on the light emission of Vibrio fischeri (luminescent bacteria test), part 3: method using freeze-dried bacteria. Dogliotti L, Hayon E (1967) Flash photolysis of persulfate ions in aqueous solutions. The sulfate and ozonide radical anions. J Phys Chem 71:2511–2516 Furman OS, Teel AL, Watts RJ (2010) Mechanism of base activation of persulfate. Environ Sci Technol 44:6423–6428 Garg S, Tripathi M, Singh S, Singh A (2013) Pentachlorophenol dechlorination and simultaneous Cr6+ reduction by Pseudomonas putida SKG-1 MTCC (10510): characterization of PCP dechlorination products, bacterial structure, and functional groups. Environ Sci Pollut Res 20:2288–2304 Hayward K (1998) Drinking water contaminant hit-list for US EPA. vol 21 House DA (1962) Kinetics and mechanism of oxidations by peroxydisulfate. Chem Rev 62:185–203 Huang K-C, Couttenye RA, Hoag GE (2002) Kinetics of heat-assisted persulfate oxidation of methyl tert-butyl ether (MTBE). Chemosphere 49:413–420 Jia H, Gu C, Li H, Fan X, Li S, Wang C (2012) Effect of groundwater geochemistry on pentachlorophenol remediation by smectitetemplated nanosized Pd0/Fe0. Environ Sci Pollut Res 19:3498–3505 Kaiser KLE, Ribo JM (1988) Photobacterium phosphoreum toxicity bioassay. II. Toxicity data compilation. Toxic Assess 3:195–237 Kim Y-H, Carraway ER (2000) Dechlorination of pentachlorophenol by zero valent iron and modified zero valent irons. Environ Sci Technol 34:2014–2017 Kolthoff IM, Miller IK (1951) The chemistry of persulfate I. The kinetics and mechanism of the decomposition of the persulfate ion in aqueous medium. J Am Chem Soc 73:3055–3059 Kuang J, Huang J, Wang B, Cao Q, Deng S, Yu G (2013) Ozonation of trimethoprim in aqueous solution: identification of reaction products and their toxicity. Water Res 47:2863–2872 Lee Y-C, Lo S-L, Chiueh P-T, Chang D-G (2009) Efficient decomposition of perfluorocarboxylic acids in aqueous solution using microwave-induced persulfate. Water Res 43:2811–2816 Liang C, Su H-W (2009) Identification of sulfate and hydroxyl radicals in thermally activated persulfate. Ind Eng Chem Res 48:5558–5562 Liang C, Wang Z-S, Bruell CJ (2007) Influence of pH on persulfate oxidation of TCE at ambient temperatures. Chemosphere 66:106–113 Lipczynska-Kochany E, Sprah G, Harms S (1995) Influence of some groundwater and surface waters constituents on the degradation of 4-chlorophenol by the Fenton reaction. Chemosphere 30:9–20 Liu X, Quan X, Bo L, Chen S, Zhao Y (2004) Simultaneous pentachlorophenol decomposition and granular activated carbon regeneration assisted by microwave irradiation. Carbon 42:415–422 Maruthamuthu P, Neta P (1978) Phosphate radicals. Spectra, acid–base equilibriums, and reactions with inorganic compounds. J Phys Chem 82:710–713 McElroy WJ, Waygood SJ (1990) Kinetics of the reactions of the SO4− radical with SO4−, S2O82−, H2O and Fe2+. J Chem Soc Faraday Trans 86:2557–2564 Nie M, Yang Y, Zhang Z, Yan C, Wang X, Li H, Dong W (2014) Degradation of chloramphenicol by thermally activated persulfate in aqueous solution. Chem Eng J 246:373–382

Niu J, Bao Y, Li Y, Chai Z (2013) Electrochemical mineralization of pentachlorophenol (PCP) by Ti/SnO2–Sb electrodes. Chemosphere 92:1571–1577 Peyton GR (1993) The free-radical chemistry of persulfate-based total organic carbon analyzers. Mar Chem 41:91–103 Pu X, Cutright T (2007) Degradation of pentachlorophenol by pure and mixed cultures in two different soils. Environ Sci Pollut Res 14:244– 250 Qi C, Liu X, Lin C, Zhang X, Ma J, Tan H, Ye W (2014) Degradation of sulfamethoxazole by microwave-activated persulfate: kinetics, mechanism and acute toxicity. Chem Eng J 249:6–14 Rodgers JD, Jedral W, Bunce NJ (1999) Electrochemical oxidation of chlorinated phenols. Environ Sci Technol 33:1453–1457 Sung M, Lee SZ, Chan HL (2012) Kinetic modeling of ring byproducts during ozonation of pentachlorophenol in water. Sep Purif Technol 84:125–131 ThanhThuy TT, Feng H, Cai Q (2013) Photocatalytic degradation of pentachlorophenol on ZnSe/TiO2 supported by photo-Fenton system. Chem Eng J 223:379–387 Tsitonaki A, Petri B, Crimi M, MosbÆK H, Siegrist RL, Bjerg PL (2010) In situ chemical oxidation of contaminated soil and groundwater using persulfate: a review. Criti Rev Env Sci Technol 40:55–91 Vallejo M, San Román MF, Ortiz I (2013) Quantitative assessment of the formation of polychlorinated derivatives, PCDD/Fs, in the electrochemical oxidation of 2-chlorophenol as function of the electrolyte type. Environ Sci Technol 47:12400–12408 Vijayalakshmi SP, Chakraborty J, Madras G (2005) Thermal and microwave-assisted oxidative degradation of poly(ethylene oxide). J Appl Polym Sci 96:2090–2096 Waldemer RH, Tratnyek PG, Johnson RL, Nurmi JT (2007) Oxidation of chlorinated ethenes by heat-activated persulfate: kinetics and products. Environ Sci Technol 41:1010–1015 WHO (2011) Guidelines for drinking-water quality, 4th edition. Malta Xu L, Yuan R, Guo Y, Xiao D, Cao Y, Wang Z, Liu J (2013) Sulfate radical-induced degradation of 2,4,6-trichlorophenol: a de novo formation of chlorinated compounds. Chem Eng J 217:169–173 Xu H-b, D-y Z, Y-j L, P-y L, C-x D (2014) Enhanced degradation of ortho-nitrochlorobenzene by the combined system of zero-valent iron reduction and persulfate oxidation in soils. Environ Sci Pollut Res 21:5132–5140 Yalfani MS, Georgi A, Contreras S, Medina F, Kopinke F-D (2011) Chlorophenol degradation using a one-pot reduction–oxidation process. Appl Catal B Environ 104:161–168 Yang S, Wang P, Yang X, Wei G, Zhang W, Shan L (2009) A novel advanced oxidation process to degrade organic pollutants in wastewater: microwave-activated persulfate oxidation. J Environ Sci 21: 1175–1180 Zhao J, Zhang Y, Quan X, Chen S (2010) Enhanced oxidation of 4chlorophenol using sulfate radicals generated from zero-valent iron and peroxydisulfate at ambient temperature. Sep Purif Technol 71: 302–307 Zhao L, Hou H, Fujii A, Hosomi M, Li F (2014) Degradation of 1,4dioxane in water with heat- and Fe2+-activated persulfate oxidation. Environ Sci Pollut Res 21:7457–7465 Zheng W, Wang X, Yu H, Tao X, Zhou Y, Qu W (2011) Global trends and diversity in pentachlorophenol levels in the environment and in humans: a meta-analysis. Environ Sci Technol 45:4668–4675 Zimbron JA, Reardon KF (2009) Fenton’s oxidation of pentachlorophenol. Water Res 43:1831–1840