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peroxymonosulfate system
Accepted Manuscript Kinetics and mechanism of sulfate radical- and hydroxyl radical-induced degradation of highly chlorinated pesticide lindane in UV/peroxymonosulfate system Sanaullah Khan, Xuexiang He, Javed Ali Khan, Hasan M. Khan, Dominic L. Boccelli, Dionysios D. Dionysiou PII: DOI: Reference:
S1385-8947(16)30807-5 http://dx.doi.org/10.1016/j.cej.2016.05.150 CEJ 15310
To appear in:
Chemical Engineering Journal
Please cite this article as: S. Khan, X. He, J.A. Khan, H.M. Khan, D.L. Boccelli, D.D. Dionysiou, Kinetics and mechanism of sulfate radical- and hydroxyl radical-induced degradation of highly chlorinated pesticide lindane in UV/peroxymonosulfate system, Chemical Engineering Journal (2016), doi: http://dx.doi.org/10.1016/j.cej. 2016.05.150
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Kinetics and mechanism of sulfate radical- and hydroxyl radicalinduced degradation of highly chlorinated pesticide lindane in UV/peroxymonosulfate system
Sanaullah Khan1, 2, Xuexiang He2, Javed Ali Khan1, 2, Hasan M. Khan1, Dominic L. Boccelli2, Dionysios D. Dionysiou2, 3, *
1
Radiation and Environmental Chemistry Laboratories, National Centre of Excellence in Physical Chemistry, University of Peshawar, Peshawar 25120, Pakistan
2
Environmental Engineering and Science Program, University of Cincinnati, Cincinnati, Ohio 45221-0012, USA 3
Nireas-International Water Research Centre, University of Cyprus, Nicosia 1678, Cyprus *
Corresponding author Email: [email protected] Tel: +1-513-556-0724; Fax: +1-513-556-2599
1
Abstract Lindane is a highly persistent chlorinated pesticide and a potent endocrine disruptor. The strong electron withdrawing property of the chlorine atoms results in a relatively low reactivity of lindane with •OH in conventional advanced oxidation processes (AOPs). In this study, the degradation of lindane by UV (254 nm)/peroxymonosulfate (UV/PMS), which can generate both •
OH and SO4•−, was investigated. A second-order rate constant of 1.3 × 109 M−1 s−1 between
lindane and SO4•− was determined using competition kinetics, suggesting a strong role of SO4•−. The degree of degradation changed with different initial solution pH, achieving 86, 92 and 55% removal of lindane at pH 4.0, 5.8 and 8.0, respectively, in 180 min, attributable to the varying concentrations of •OH and SO4•−. The addition of common water quality constituents, e.g., humic acid or inorganic anions, at pH 5.8 showed a varied inhibition effect with the 61% degradation in the presence of 1.0 mg L−1 humic acid, and 45, 60, 88 and 91% degradation in the presence of 1 mM CO32−, HCO3−, SO42− and Cl−, respectively, in 180 min. With the kinetics being demonstrated to be feasible, the degradation mechanism of lindane by UV/PMS was also assessed. Based on the detected by-products through GC-MS analysis, plausible reaction pathways were proposed, suggesting dechlorination, chlorination, dehydrogenation and hydroxylation via •OH and/or SO4•− attack. Meanwhile, reasonable mineralization efficiency was observed, with a 56% total organic carbon removal in 360 min, at an initial PMS concentration of 500 µM. Results from both degradation kinetics and transformation mechanisms indicate that UV/PMS is a potential method for the treatment of water contaminated with lindane.
Keywords: Lindane; UV/peroxymonosulfate; Water quality parameters; Second-order rate constant; Degradation mechanism. 2
1. Introduction The extensive use of synthetic pesticides for diverse agricultural and non-agricultural purposes, though succeeding to a large extent in high crop yields and certain other goals, such as vector control, has resulted in the occurrence of pesticides in various environmental segments around the world. Pesticide contamination of surface waters has been well documented, constituting a major issue of concerns at local, regional, national, and global scale [1, 2]. Organochlorine pesticides represent one of the most toxic and persistent classes of the synthetic pesticides and are a significant source of endocrine disrupting chemicals (CDCs) causing an adverse effect on human health and the environment [3]. Lindane ((1r,2R,3S,4r,5R,6S)-hexachlorocyclohexane, also called γ-HCH) is one of the most widely studied organochlorine pesticides in the last few decades [4-10]. It has recently been included in Annex A of the Stockholm Convention POP (persistent organic pollutant) list by the POP review committee [4]. Since 1940s, lindane has been widely used for agricultural pest control, seed treatments, poultry and livestock treatment, household vector control, lumber protection, lice and scabies treatment, and even in rodent baits treatment [5, 6]. The total global usage of lindane was estimated to be 720,000 metric tons by 1995 [7]. Lindane is currently being used in the United States and Canada for seed treatment [8]. In addition to the large amount of lindane employed, thousands of tons of unused stockpiles are stored in containers waiting for a further disposal [9]. Owing to its global usage, long-range transportation, bioaccumulation, and high persistency, lindane appears as a ubiquitous contaminant in the environment [10]. There have been increasing research studies into the removal of lindane from the contaminated water, including the development of new and effective technologies.
3
Advanced oxidation processes (AOPs) based on the generation of strong oxidizing and highly reactive species such as hydroxyl radical (•OH) and/or sulfate radical (SO4•−) have proved to be promising alternatives to conventional treatment processes for the treatment of water and wastewater organic pollutants [11]. Hydroxyl radical based AOPs (•OH-based AOPs), such as titanium dioxide (TiO2) and/or polyoxometalates (POMs) photocatalysis, UV/H2O2, and photoFenton process, have been previously studied for the degradation of lindane [12-15]. Sulfate radical based AOPs (SO4•−-based AOPs) are comparatively a new group of technologies, which require more investigations on their potential applicability in treating the contaminated water [16-21]. SO4•− can be generated by the activation of persulfate (PS) and/or peroxymonosulfate (PMS) with transition metals [17], elevated temperature or pH [18, 19], and/or UV irradiation [20]. Previous research studies have demonstrated that SO4•−-based AOPs employing PS as a precursor oxidant can effectively degrade lindane in water and soil [22, 23]. However, there is still very limited study on the degradation of lindane by SO4•−-based AOPs employing PMS [24]. Peroxymonosulfate (available as a triple potassium salt with a commercial name of Oxone®, 2KHSO5· KHSO4·K2SO4) is a highly versatile and an environmentally friendly oxidant [25, 26]. It has received great attention and application in water disinfection and decontamination [19-21]. Different from the studied activation of PMS by transition metals such as Co(II) and iron [17, 24], UV (namely UV-254 nm) is chosen in this study to activate PMS, an established process that can generate both SO4•− and •OH as shown below in reaction (1) [20]. + ℎ → • + • ( = 1.04)
[27]
(1)
The main objective of this study was to investigate the efficiency of UV/PMS process for the degradation of lindane in synthetic water. The influence of water quality and process parameters, such as pH, presence of natural organic matter (NOM) and common inorganic anions, and initial 4
concentrations of pollutant and oxidant, was evaluated. To better compare the fundamental •OH and SO4•− reactions, the second-order rate constant of lindane with SO4•− was determined for the first time in this study using a competition kinetics approach. Transformation by-products were detected using GC-MS and a potential reaction pathway was subsequently proposed. Lastly, mineralization of lindane by UV/PMS process was also assessed. The obtained data could provide useful information on the applicability of UV/PMS based AOPs for the treatment of water contaminated with lindane or other chlorinated pesticides. 2. Materials and methods 2.1 Materials Lindane (C6H6Cl6, 97%) and Oxone® were obtained from Sigma–Aldrich (St. Louis, MO, USA). All other chemicals, such as sodium sulfate (Na2SO4), sodium chloride (NaCl), sodium carbonate (Na2CO3) and sodium bicarbonate (NaHCO3), were of high purity and obtained from Fisher Scientific (Pittsburgh, PA, USA). Standard Suwannee River humic acid (SRHA) was purchased from the International Humic Substances Society (IHSS, University of Minnesota, St. Paul, MN, USA) and used as a representative of NOM. The initial pH of the solution was adjusted by 0.1 N HCl or NaOH. All solutions were prepared using Milli-Q water (18.2 MΩ·cm, Milli-pore Corp., Billerica, MA). All chemicals were used as received. 2.2 Analytical Methods Gas chromatograph (GC, Agilent 6890) with a mass selective detector (Agilent 5975, Wilmington, DE, USA) was used for the quantification of lindane and detection of its reaction by-products. Detailed analytical parameters were described in our previous study [24]. Briefly, sample extraction was performed using a solid phase micro extraction (SPME) technique. Separation of the analytes was achieved on an HP-5MS (5% phenyl methylsiloxane) capillary 5
column (30 m, i.d. 0.25 µm). Mass spectra, recorded from m/z 50-550, were obtained in an electron impact ionization mode (EI+) at 70 eV. The data were analyzed using a mass spectral search program (NIST, USA) installed in the GC-MS and the obtained spectra were compared with those of the standards in the NIST library. Determination of total organic carbon (TOC), as non-purgeable organic carbon, was performed using a Shimadzu VCSH-ASI TOC analyzer. 2.3 Degradation experiments Experiments were performed in a Pyrex Petri dish with a quartz cover. Total solution volume was 10 mL. The radiation source consisted of two 15W low-pressure Hg UV lamps (Cole– Parmer) which emit light primarily at λ = 254 nm. The average fluence rate of the UV lamps was found to be 0.10 mW/cm2 as determined by three different methods: ferrioxalate actinometry [28], iodide/iodate actinometry [29] and a calibrated digital radiometer (Model IL 1700, XRD (XRL) 140T254 low profile germicidal probe, International Light, Co., Newburyport, MA). The experiments were conducted in triplicate unless stated otherwise. The error bars in the figures represent the standard error of the mean. 2.4 Determination of second-order rate constant Competition kinetics studies were conducted to determine the second-order rate constant between lindane and SO4•−. The reference compound used was meta-toluic acid (m-TA), which has a known second-order rate constant of 2.0 × 109 M−1 s−1 with SO4•− [30]. The initial concentration of lindane, m-TA, PS, tert-butanol and phosphate buffer (pH = 5.8) were 3.43 µM, 3.43 µM, 250 µM, 50 mM and 5 mM, respectively. The UV direct photolysis was also considered in the rate constant calculation, as shown in reaction (2) [31].
6
k
•−
SO4
(s)
k − k s -UV = s -UV / PS k ref -UV / PS − kref -UV
k •− SO4 ( ref )
(2)
where kSO4 •−, kUV/PS and kUV represent the second-order rate constant with SO4•−, kobs in UV/PS and kobs in UV only condition, while “s” and “ref” stand for the substrate and reference compound, respectively. 3. Results and discussion 3.1 Degradation of lindane by UV/PMS process As shown in Figure 1, neither direct UV photolysis nor PMS alone showed an apparent degradation of lindane during 180 min reaction. However, the incorporation of UV and PMS promoted significantly the degradation efficiency, with a kobs of 1.42 × 10−2 min−1 and a 92% lindane removal in 180 min, which was attributed to the reactive radical species generated following reaction (1). The obtained results are in good agreement with the literature data on the removal of other organic pollutants from water and wastewater using a UV/PMS system [32, 33]. The •OH is a non-selective radical that can react with organic compounds mainly through hydrogen abstraction or hydroxyl additions with a second-order rate constant of 108 – 1010 M−1 s−1 [34]. Due to the presence of strong electron withdrawing chlorine atoms, the second-order rate constant of lindane with •OH has been reported in literature to be in a low range of 5.2 × 108 to 1.1 × 109 M−1 s−1 [35]. The SO4•−, though also highly reactive, exhibits a higher selectivity [36]. Its reaction with electron deficit compounds such as perfluorooctanoic acids (PFOAs) is generally faster than •OH [37, 38]. The second-order rate constant of lindane with SO4•− was subsequently determined in this study to be 1.3 × 109 M−1 s−1, consistent with the above hypothesis.
7
3.2 Influence of process parameters (i.e., initial concentrations of PMS or lindane) The •OH and SO4•− can recombine easily, react with non-target species including solvent molecules and by-products, and be converted into unreactive species, which can subsequently decrease the reaction of the radicals with the target compound. Major chemical reactions involved in a PMS-based system are shown in reactions (3)-(10) [34, 36, 39]. Initial concentrations of PMS or lindane may have influences on the efficiency of UV/PMS process by way of influencing those reactions. As shown in Figure 2, the removal efficiency of lindane was found to increase significantly with increasing the initial concentrations of PMS, with a percentage removal of 58, 92, 98 and 100% in 180 min obtained, respectively, at an initial PMS concentration of 125, 250, 500 and 1000 µM. This trend can be explained by reaction (1), with the rate of radical formation increasing with an increase in the concentration of PMS [39]. An eight-time increase in PMS concentration led to a 6.2-time rise in the value of kobs, i.e., 4.77 × 10−3 min−1 at 125 µM PMS vs 2.97 × 10−2 min−1 at 1000 µM PMS (inset of Figure 2), demonstrating an inhibition role of a higher concentration of PMS. Possible explanations for this observation include the interaction of the reactive radicals with PMS following reactions (7) and (8), and the self-combination reactions among the reactive radicals (reactions (3)-(5)) especially at higher initial concentrations of the oxidant [39]. • + • → ( = 4 × 10 )
[40]
(3)
•
+ • → ( = (0.95 ± 0.08) × 10 )
[41]
(4)
•
+ • → ( = 5.3 × 10" )
[42]
(5)
• + → # + + •
[36]
(6)
• + → • + + # ( < 1.0 × 10 )
[43]
(7)
[36]
(8)
•
+ → • +
8
• + → + • ( = 8.3 × 10% , ' > 11) •
+ → • + ( = 1.2 × 10 , *+*+,-. ')
[36]
(9)
[34]
(10)
As seen in the Figure 3 and Table 1, with an increasing lindane concentration from 0.68 to 3.43 and 6.86 µM, the percentage degradation of lindane decreased from 100% in 120 min to 92 and 80%, respectively, in 180 min; while the kobs decreased from 3.51 × 10−2 to 1.42 × 10−2 and 7.89 × 10−3 min−1, respectively. The corresponding half-life values of these reactions were calculated and found to be 19.7, 48.8, and 87.8 min. A plausible reason could be from the formation of a large number of by-products (as will be discussed below on Figure 5) which competed with lindane for the reactive radicals and the concentration of which increased with increasing the initial concentration of lindane [44]. 3.3 Influence of water quality parameters (i.e., pH, humic acid or inorganic anions) Reactive radicals can react with both organic and inorganic compounds. Common water quality parameters, such as pH, inorganic anions and humic acids, may therefore compete with target compound for radical species leading to positive or negative impacts on the degradation kinetics [45]. Therefore, it is important to consider and evaluate the impact of these variables on the removal efficiency of lindane by UV/PMS. The SO4•− reactions such as those shown in reactions (6) and (7) could lead to the generation of hydronium ions (H3O+) which could subsequently result in a decrease in solution pH, as measured by the pH of 3.8, 5.4 and 7.2 for the initial solution pH of 4.0, 5.8 and 8.0, respectively, after 180 min. As shown in Figure 4, the removal efficiency was affected by varying the initial solution pH, with 86, 92 and 55% removal of lindane at initial pH 4.0, 5.8 and 8.0, respectively, in 180 min. The slight difference in the removal efficiency (~6%) by changing the pH from 4.0 to 5.8 was probably due to the unchanged radical quantum yield during PMS 9
photolysis in this pH range [27]. The decreased removal efficiency at higher solution pH, on the other hand, could be attributed to the scavenging of SO4•− and •OH by HO− in accordance with reactions (9) and (10), respectively [34, 36]. The effect of humic acid on the degradation of lindane by UV/PMS process is also illustrated in Figure 4. The addition of 1.0 mg L−1 humic acid decreased greatly the removal efficiency of UV/PMS, i.e., from 92 to 61% after 180 min, which could result from the scavenging effect of NOM towards •OH and SO4•− via reactions (11) and (12)), respectively [46, 47]. This result is consistent with the literature reports showing that efficiency of the UV/PMS system for atrazine removal decreased from 98% to 23% when 3.2 mg L−1 of humic acid was added to the reaction mixture [48]. Therefore, when applying UV/PMS AOPs, a pre-treatment aiming to remove organic matter may be considered in order to minimize the negative effects from those non-target compounds. The effect of inorganic anions on the efficiency of UV/PMS process was also investigated. The anions selected included CO32−, HCO3−, SO42− and Cl−, the presence of which led to different degrees of inhibition on the percentage removal of lindane, as shown in Figure 4. The presence of HCO3− or CO32− caused a significant inhibitory effect, attributable to the relatively strong scavenging of reactive radicals by these anions (reactions (13) – (16)) [34, 39]. The higher reactivity of CO32− with SO4•− and •OH than HCO3− allowed the former to show a higher inhibitory effect on the removal of lindane. Cl− can have a dual role on the efficiency of photochemical degradation processes, attributed to the transformation of SO4 •− and •OH into reactive but strongly selective radicals, e.g., Cl•, Cl2•− and ClOH•−, inducing either a promoting or an inhibiting effect on the removal of the organics by activated persulfate or peroxymonosulfate [49, 50]. Some of the important aqueous phase reactions of Cl• and Cl2 •− are shown in reactions 10
(17)-(27) [51, 52]. Thus despite the high reactivity of Cl− towards •OH and SO4•− (reactions (17) and (18)) [49], a negligibly small adverse effect was generally observed in this study on the efficiency of UV/PMS process. The overall removal efficiency of the UV/PMS process at an initial pH of 5.8 decreased in the following order: control ~ SO42− ~ Cl− > humic acid ~ HCO3− > CO32−. •
+ / → 0123456 ( = 2.23 × 10 7 (82+ 9 ) )
[46]
(11)
• + / → 0123456 ( > 6 × 10; 7 (82+ 9) )
[47]
(12)
9< + • → 9< • + ( = 4.1 × 10; )
[53]
(13)
9< + • → 9< • + + # ( = 3.5 × 10; )
[54]
(14)
9< + • → + 9< • ( = 3.9 × 10 )
[34]
(15)
9< + • → + 9< • ( = 8.5 × 10; )
[34]
(16)
9+ + • → 9+ • + ( = 2.6 × 10 )
[53]
(17)
9+ + • → 9+• ( = 4.3 × 10" )
[34]
(18)
9+• → 9+ + • ( = 6.1 × 10" )
[51]
(19)
9+ • + 9+ → 9+ • ( = 8.5 × 10" )
[52]
(20)
9+ • + 9+ • → 9+
[52]
(21)
9+ • + 9+ • → 9+ + 9+
[52]
(22)
9+ • + 9+ • → 9+ + 29+
[52]
(23)
9+ • + → 9+ • (2.5 × 10 )
[52]
(24)
9+ • + → 9+ • + 9+ (1.3 × 10< )
[52]
(25)
9+ • → 9+ • + 9+ (6.0 × 10 )
[51]
(26)
9+ • + → 9+ • + 9+ (4.0 × 10; )
[51]
(27) 11
3.4 By-product analysis and potential reaction mechanism With the kinetics being demonstrated to be feasible, the mechanism of the degradation of lindane by UV/PMS was assessed and presented in this section. There were mainly six reaction by-products identified by GC-MS in this study: 1,1,2,3,4,5,6-heptachlorocyclohexane (HeCH), 1,2,3,4,5,6-hexachlorobenzene (HCB), 1,3,4,5,6-pentachlorocyclohexene (PCCH), 3,4,5,6tetrachlorocyclohexene (TeCCH), 1,2,4-trichlorobenzene (TCB) and 2,4,5-trichlorophenol (TCP). These by-products have been reported previously in various oxidative studies on lindane, e.g., HeCH, HCB, PCCH, TeCCH and TCP in POMs photocatalysis [13], TCB and HCB in the photo-Fenton reaction [15], as well as HeCH, PCCH and TeCCH in TiO2 photocatalysis [55]. In this study, the exact reacting radical could not be distinguished with certainty because of the coexistence of both •OH and SO4•− in the reaction solution [31, 56]. Due to the similarities in the reaction mechanism of these two radicals, it is very likely that the detected by-products can be from either radical reaction. A potential reaction pathway for the degradation of lindane was proposed and is shown in Figure 5, including dechlorination, chlorination, dehydrogenation and hydroxylation. Dechlorination was presumably resulted from homolytic scission of the C-Cl bond upon UV excitation [57, 58]. As a result, chlorine radical (Cl•) was released from lindane, leaving a carbon centered radical. A subsequent abstraction of hydrogen from an adjoining carbon atom by •
OH and/or SO4•− might lead to the formation of PCCH. A further loss of chlorine resulted in the
formation of lesser chlorinated by-products such as TeCCH and TCB. Chlorination of lindane could result from its reaction with chlorine radical (Cl•) or dichlorine radical anion (Cl2•−). Cl• and Cl2•− are strong oxidizing species with E0 of 2.4 and 2.0 V, respectively [51]. These oxidizing species may react with organic compounds via addition to 12
double bond, hydrogen abstraction or electron-transfer reactions [59]. Thus an abstraction of hydrogen from lindane via Cl• led to the formation of HeCH. A similar explanation has been provided by Antonaraki et al. [13] employing POMs photocatalysis of lindane. In fact, the interaction of Cl− with SO4•− with the subsequent formation of the chlorinated organic compounds in water has been commonly reported in literature [60]. It has been widely reported that chlorine group was mainly responsible for the toxicity of chlorinated organic compounds [61]. The dechlorination achieved in water treatment processes may therefore result in a reduction in the overall toxicity [61]. However, with the chlorination to be one of the most significant reaction pathways in the transformation of lindane in UV/PMS, whether the toxicity of the reaction solution was reduced needs more investigations. Dehydrogenation might occur with the abstraction of two adjoining hydrogen atoms by •
OH and/or SO4•− attack. The resulting conjugated double bonded HCB by-product with more
thermal stability allowed the dehydrogenation to be kinetically feasible [62]. The formation of HCB from lindane via hydrogen abstraction by •OH was also reported elsewhere [13]. Hydroxylation is a process that introduces a hydroxyl group (-OH) into an organic compound. The addition of the electrophilic •OH to TCB formed a carbon centered radical, which by addition of O2 yielded a peroxy radical. After releasing HO2 •, TCP could be formed (reaction (28)). Hydroxylation of chlorobenzene and other chlorinated aromatic compound through such a pathway has been proposed earlier [63, 64].
(28)
13
In SO4•− mediated mechanisms, SO4•− oxidized the aromatic ring to a radical cation, which upon hydrolysis led to the formation of hydroxycyclohexadienyl radical. The resulting radical, after reaction with O2 and releasing subsequently HO2•, was converted into a hydroxylated phenolic by-product (reaction (29)) [30, 60, 65]. It should be noted, however, there were different positions available at the ring for hydroxylation reaction, which could be influenced by both steric hindrance and resonance stability [56, 66]. More studies are needed to differentiate the exact structure of TCP.
(29) Though not identified by our method, ring opening and cleavage by-products were also expected to form. The intermediate by-products HCB and TCP, for example, were known to mineralize into CO2, H2O and Cl− with an extended reaction time by photocatalytic and photochemical transformations [67, 68]. 3.5 Mineralization study As shown in the inset of Figure 6, though much less effective than the removal of the parent compound, a 56% TOC removal (TOC0 = 1.24 mg/L) by UV/PMS in 360 min could still be obtained. Increasing initial PMS concentration had an enhancing effect on lindane mineralization, with a 36, 49, 56 and 60% TOC removal achieved at an initial PMS concentration of 125, 250, 500, and 1000 µM, respectively, in 360 min (Figure 6). Ahmed and Chiron [33] also observed a similar mineralization trend during the oxidation of ciprofloxacin by UV/PMS. This could again be attributed to the increased concentration of •OH and/or SO4•− with 14
an increasing concentration of PMS, consistent with the effect of PMS concentrations on the degradation of lindane. 4 Conclusions In this study, the degradation of the environmentally persistent chlorinated pesticide lindane was investigated both kinetically and mechanistically. Compared to common hydroxyl radical reaction, SO4•− exhibited a slightly higher reactivity towards lindane as determined from its second-order rate constant of 1.3 × 109 M−1 s−1. Process parameters such as initial concentrations of PMS or lindane, as well as common water quality parameters such as organic humic acid or inorganic anions, showed a varied degree of impacts on the observed pseudo firstorder rate constant (kobs). The presence of humic acid, CO32− or HCO3− caused a strong inhibiting effect while the presence of SO42− or Cl− exerted a negligible effect on the efficiency of UV/PMS process. Various degradation by-products as identified by GC-MS revealed dechlorination, chlorination, dehydrogenation and hydroxylation to be potential transformation steps. Ring opening and cleavage could also be achieved as demonstrated indirectly by the significant decrease in TOC. This study shows UV/PMS based AOPs to be an effective method for the removal of lindane from aqueous environment. Acknowledgments The Higher Education Commission (HEC), Islamabad, Pakistan is highly acknowledged for funding this research project through an International Research Support Initiative Program (IRSIP). This work was also partially funded by the Cyprus Research Promotion Foundation through Desmi 2009-2010 which is co-funded by the Republic of Cyprus and the European Regional
Development
Fund
of
the
EU
under
contract
number NEA
IPODOMI/STRATH/0308/09. 15
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24
Figure Captions Figure 1. Degradation of lindane by UV/PMS. [Lindane]0 = 3.43 µM, [PMS]0 = 250 µM, pH = 5.8. Figure 2. Effect of initial concentrations of PMS on the efficiency of UV/PMS process. Inset: variation of kobs (min−1) with [PMS]0. [Lindane]0 = 3.43 µM, pH = 5.8. Figure 3. Effect of initial concentrations of lindane on the efficiency of UV/PMS process. [Lindane]0 = 0.68, 3.43, and 6.86 µM, [PMS]0 = 250 µM, pH = 5.8. Figure 4. Effect of initial pH, humic acid or inorganic anions on the percentage removal of lindane in 180 min. [Lindane]0 = 3.43 µM, [PMS]0 = 250 µM, [humic acid]0 = 1.0 mg L−1, [inorganic anions]0 = 1000 µM, pH = 5.8 except noted. Figure 5. Proposed degradation pathway of lindane by UV/PMS. [Lindane]0 = 17.15 µM, [PMS]0 = 500 µM, pH = 5.8. Figure 6. Plots of TOC/TOC0 vs time by UV/PMS at using different initial concentrations of PMS, i.e., [PMS]0 = 125-1000µM, TOC0 = 1.24 mg/L. Inset: TOC vs lindane degradation : [Lindane]0 = 17.15 µM, [PMS]0 = 500 µM, pH = 5.8.
Table Captions Table 1. Pseudo first-order rate constant (kobs) and half-life (t1/2) at different initial concentrations of lindane in UV/PMS process.
25
Figure 1.
1.0
C/C0
0.8
PMS only UV photolysis UV/PMS
0.6
0.4
0.2
0.0 0
50
100
150
200
Time (min)
26
Figure 2.
[PMS]0 = 125 µM 1.0
[PMS]0 = 500 µM [PMS]0 = 1000 µM
0.8
kobs (min-1)
0.03
[PMS]0 = 250 µM
0.02 0.01 0.00
C/C0
0
200 400 600 800 1000
0.6
[PMS]0 (µM)
0.4
0.2
0.0 0
50
100
150
200
Time (min)
27
Figure 3.
[Lindane]0 = 6.86 mM [Lindane]0 = 3.43 mM [Lindane]0 = 0.68 mM
Lindane concentration (µM)
8
6
4
2
0
0
50
100
150
200
Time (min)
28
Figure 4.
Removal efficiency (%)
100
80
60
40
20
0 Humic acid CO32-
-
HCO3
Cl-
SO42-
Control pH = 4.0 pH = 8.0 (pH = 5.8)
Effect of water quality parameters
29
Figure 5.
30
Removal efficiency
Figure 6.
TOC/TOC0
1.0
0.8
1.0
TOC Lindane
0.8 0.6 0.4 0.2 0.0 0
100
200
300
400
Time (min)
0.6 125 µM 250 µM 500 µM 1000 µM
0.4
0.2 0
100
200
300
400
Time (min)
31
Table 1. ______________________________________________________________________________ [lindane]0 (µM)
Rate constant (min−1)
t1/2 (min)
R2
0.68
3.51 × 10−2
19.7
0.992
3.43
1.42 × 10−2
48.8
0.987
6.68
7.89 × 10−3
87.8
0.981
32
Graphical abstract
Lindane removal efficiency (%)
100
80
60
40
20
0
M NO
m (1
) g/L
2-
3 CO
(
M) 1m
- ( O3 C H
M) 1m
- (1 Cl
) mM
2-
4 SO
) ) ) Cl OH M) 5.8 H a ( m H (N .0 (p (1 .0 ol =4 8 r t = n pH Co pH
Effecet of water quality parameters
33
Highlights
•
Removal of lindane by UV-C/peroxymonosulfate based AOPs was investigated.
•
Lindane showed a comparable or slightly higher reactivity toward SO4•− than HO•.
•
Degradation efficiency was significantly affected in presence of NOM or alkalinity.
•
A reasonable mineralization of lindane was achieved by UV/PMS.
•
Transformation mechanism was proposed based on by-products identified via GC-MS.
34
S1385-8947(16)30807-5 http://dx.doi.org/10.1016/j.cej.2016.05.150 CEJ 15310
To appear in:
Chemical Engineering Journal
Please cite this article as: S. Khan, X. He, J.A. Khan, H.M. Khan, D.L. Boccelli, D.D. Dionysiou, Kinetics and mechanism of sulfate radical- and hydroxyl radical-induced degradation of highly chlorinated pesticide lindane in UV/peroxymonosulfate system, Chemical Engineering Journal (2016), doi: http://dx.doi.org/10.1016/j.cej. 2016.05.150
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Kinetics and mechanism of sulfate radical- and hydroxyl radicalinduced degradation of highly chlorinated pesticide lindane in UV/peroxymonosulfate system
Sanaullah Khan1, 2, Xuexiang He2, Javed Ali Khan1, 2, Hasan M. Khan1, Dominic L. Boccelli2, Dionysios D. Dionysiou2, 3, *
1
Radiation and Environmental Chemistry Laboratories, National Centre of Excellence in Physical Chemistry, University of Peshawar, Peshawar 25120, Pakistan
2
Environmental Engineering and Science Program, University of Cincinnati, Cincinnati, Ohio 45221-0012, USA 3
Nireas-International Water Research Centre, University of Cyprus, Nicosia 1678, Cyprus *
Corresponding author Email: [email protected] Tel: +1-513-556-0724; Fax: +1-513-556-2599
1
Abstract Lindane is a highly persistent chlorinated pesticide and a potent endocrine disruptor. The strong electron withdrawing property of the chlorine atoms results in a relatively low reactivity of lindane with •OH in conventional advanced oxidation processes (AOPs). In this study, the degradation of lindane by UV (254 nm)/peroxymonosulfate (UV/PMS), which can generate both •
OH and SO4•−, was investigated. A second-order rate constant of 1.3 × 109 M−1 s−1 between
lindane and SO4•− was determined using competition kinetics, suggesting a strong role of SO4•−. The degree of degradation changed with different initial solution pH, achieving 86, 92 and 55% removal of lindane at pH 4.0, 5.8 and 8.0, respectively, in 180 min, attributable to the varying concentrations of •OH and SO4•−. The addition of common water quality constituents, e.g., humic acid or inorganic anions, at pH 5.8 showed a varied inhibition effect with the 61% degradation in the presence of 1.0 mg L−1 humic acid, and 45, 60, 88 and 91% degradation in the presence of 1 mM CO32−, HCO3−, SO42− and Cl−, respectively, in 180 min. With the kinetics being demonstrated to be feasible, the degradation mechanism of lindane by UV/PMS was also assessed. Based on the detected by-products through GC-MS analysis, plausible reaction pathways were proposed, suggesting dechlorination, chlorination, dehydrogenation and hydroxylation via •OH and/or SO4•− attack. Meanwhile, reasonable mineralization efficiency was observed, with a 56% total organic carbon removal in 360 min, at an initial PMS concentration of 500 µM. Results from both degradation kinetics and transformation mechanisms indicate that UV/PMS is a potential method for the treatment of water contaminated with lindane.
Keywords: Lindane; UV/peroxymonosulfate; Water quality parameters; Second-order rate constant; Degradation mechanism. 2
1. Introduction The extensive use of synthetic pesticides for diverse agricultural and non-agricultural purposes, though succeeding to a large extent in high crop yields and certain other goals, such as vector control, has resulted in the occurrence of pesticides in various environmental segments around the world. Pesticide contamination of surface waters has been well documented, constituting a major issue of concerns at local, regional, national, and global scale [1, 2]. Organochlorine pesticides represent one of the most toxic and persistent classes of the synthetic pesticides and are a significant source of endocrine disrupting chemicals (CDCs) causing an adverse effect on human health and the environment [3]. Lindane ((1r,2R,3S,4r,5R,6S)-hexachlorocyclohexane, also called γ-HCH) is one of the most widely studied organochlorine pesticides in the last few decades [4-10]. It has recently been included in Annex A of the Stockholm Convention POP (persistent organic pollutant) list by the POP review committee [4]. Since 1940s, lindane has been widely used for agricultural pest control, seed treatments, poultry and livestock treatment, household vector control, lumber protection, lice and scabies treatment, and even in rodent baits treatment [5, 6]. The total global usage of lindane was estimated to be 720,000 metric tons by 1995 [7]. Lindane is currently being used in the United States and Canada for seed treatment [8]. In addition to the large amount of lindane employed, thousands of tons of unused stockpiles are stored in containers waiting for a further disposal [9]. Owing to its global usage, long-range transportation, bioaccumulation, and high persistency, lindane appears as a ubiquitous contaminant in the environment [10]. There have been increasing research studies into the removal of lindane from the contaminated water, including the development of new and effective technologies.
3
Advanced oxidation processes (AOPs) based on the generation of strong oxidizing and highly reactive species such as hydroxyl radical (•OH) and/or sulfate radical (SO4•−) have proved to be promising alternatives to conventional treatment processes for the treatment of water and wastewater organic pollutants [11]. Hydroxyl radical based AOPs (•OH-based AOPs), such as titanium dioxide (TiO2) and/or polyoxometalates (POMs) photocatalysis, UV/H2O2, and photoFenton process, have been previously studied for the degradation of lindane [12-15]. Sulfate radical based AOPs (SO4•−-based AOPs) are comparatively a new group of technologies, which require more investigations on their potential applicability in treating the contaminated water [16-21]. SO4•− can be generated by the activation of persulfate (PS) and/or peroxymonosulfate (PMS) with transition metals [17], elevated temperature or pH [18, 19], and/or UV irradiation [20]. Previous research studies have demonstrated that SO4•−-based AOPs employing PS as a precursor oxidant can effectively degrade lindane in water and soil [22, 23]. However, there is still very limited study on the degradation of lindane by SO4•−-based AOPs employing PMS [24]. Peroxymonosulfate (available as a triple potassium salt with a commercial name of Oxone®, 2KHSO5· KHSO4·K2SO4) is a highly versatile and an environmentally friendly oxidant [25, 26]. It has received great attention and application in water disinfection and decontamination [19-21]. Different from the studied activation of PMS by transition metals such as Co(II) and iron [17, 24], UV (namely UV-254 nm) is chosen in this study to activate PMS, an established process that can generate both SO4•− and •OH as shown below in reaction (1) [20]. + ℎ → • + • ( = 1.04)
[27]
(1)
The main objective of this study was to investigate the efficiency of UV/PMS process for the degradation of lindane in synthetic water. The influence of water quality and process parameters, such as pH, presence of natural organic matter (NOM) and common inorganic anions, and initial 4
concentrations of pollutant and oxidant, was evaluated. To better compare the fundamental •OH and SO4•− reactions, the second-order rate constant of lindane with SO4•− was determined for the first time in this study using a competition kinetics approach. Transformation by-products were detected using GC-MS and a potential reaction pathway was subsequently proposed. Lastly, mineralization of lindane by UV/PMS process was also assessed. The obtained data could provide useful information on the applicability of UV/PMS based AOPs for the treatment of water contaminated with lindane or other chlorinated pesticides. 2. Materials and methods 2.1 Materials Lindane (C6H6Cl6, 97%) and Oxone® were obtained from Sigma–Aldrich (St. Louis, MO, USA). All other chemicals, such as sodium sulfate (Na2SO4), sodium chloride (NaCl), sodium carbonate (Na2CO3) and sodium bicarbonate (NaHCO3), were of high purity and obtained from Fisher Scientific (Pittsburgh, PA, USA). Standard Suwannee River humic acid (SRHA) was purchased from the International Humic Substances Society (IHSS, University of Minnesota, St. Paul, MN, USA) and used as a representative of NOM. The initial pH of the solution was adjusted by 0.1 N HCl or NaOH. All solutions were prepared using Milli-Q water (18.2 MΩ·cm, Milli-pore Corp., Billerica, MA). All chemicals were used as received. 2.2 Analytical Methods Gas chromatograph (GC, Agilent 6890) with a mass selective detector (Agilent 5975, Wilmington, DE, USA) was used for the quantification of lindane and detection of its reaction by-products. Detailed analytical parameters were described in our previous study [24]. Briefly, sample extraction was performed using a solid phase micro extraction (SPME) technique. Separation of the analytes was achieved on an HP-5MS (5% phenyl methylsiloxane) capillary 5
column (30 m, i.d. 0.25 µm). Mass spectra, recorded from m/z 50-550, were obtained in an electron impact ionization mode (EI+) at 70 eV. The data were analyzed using a mass spectral search program (NIST, USA) installed in the GC-MS and the obtained spectra were compared with those of the standards in the NIST library. Determination of total organic carbon (TOC), as non-purgeable organic carbon, was performed using a Shimadzu VCSH-ASI TOC analyzer. 2.3 Degradation experiments Experiments were performed in a Pyrex Petri dish with a quartz cover. Total solution volume was 10 mL. The radiation source consisted of two 15W low-pressure Hg UV lamps (Cole– Parmer) which emit light primarily at λ = 254 nm. The average fluence rate of the UV lamps was found to be 0.10 mW/cm2 as determined by three different methods: ferrioxalate actinometry [28], iodide/iodate actinometry [29] and a calibrated digital radiometer (Model IL 1700, XRD (XRL) 140T254 low profile germicidal probe, International Light, Co., Newburyport, MA). The experiments were conducted in triplicate unless stated otherwise. The error bars in the figures represent the standard error of the mean. 2.4 Determination of second-order rate constant Competition kinetics studies were conducted to determine the second-order rate constant between lindane and SO4•−. The reference compound used was meta-toluic acid (m-TA), which has a known second-order rate constant of 2.0 × 109 M−1 s−1 with SO4•− [30]. The initial concentration of lindane, m-TA, PS, tert-butanol and phosphate buffer (pH = 5.8) were 3.43 µM, 3.43 µM, 250 µM, 50 mM and 5 mM, respectively. The UV direct photolysis was also considered in the rate constant calculation, as shown in reaction (2) [31].
6
k
•−
SO4
(s)
k − k s -UV = s -UV / PS k ref -UV / PS − kref -UV
k •− SO4 ( ref )
(2)
where kSO4 •−, kUV/PS and kUV represent the second-order rate constant with SO4•−, kobs in UV/PS and kobs in UV only condition, while “s” and “ref” stand for the substrate and reference compound, respectively. 3. Results and discussion 3.1 Degradation of lindane by UV/PMS process As shown in Figure 1, neither direct UV photolysis nor PMS alone showed an apparent degradation of lindane during 180 min reaction. However, the incorporation of UV and PMS promoted significantly the degradation efficiency, with a kobs of 1.42 × 10−2 min−1 and a 92% lindane removal in 180 min, which was attributed to the reactive radical species generated following reaction (1). The obtained results are in good agreement with the literature data on the removal of other organic pollutants from water and wastewater using a UV/PMS system [32, 33]. The •OH is a non-selective radical that can react with organic compounds mainly through hydrogen abstraction or hydroxyl additions with a second-order rate constant of 108 – 1010 M−1 s−1 [34]. Due to the presence of strong electron withdrawing chlorine atoms, the second-order rate constant of lindane with •OH has been reported in literature to be in a low range of 5.2 × 108 to 1.1 × 109 M−1 s−1 [35]. The SO4•−, though also highly reactive, exhibits a higher selectivity [36]. Its reaction with electron deficit compounds such as perfluorooctanoic acids (PFOAs) is generally faster than •OH [37, 38]. The second-order rate constant of lindane with SO4•− was subsequently determined in this study to be 1.3 × 109 M−1 s−1, consistent with the above hypothesis.
7
3.2 Influence of process parameters (i.e., initial concentrations of PMS or lindane) The •OH and SO4•− can recombine easily, react with non-target species including solvent molecules and by-products, and be converted into unreactive species, which can subsequently decrease the reaction of the radicals with the target compound. Major chemical reactions involved in a PMS-based system are shown in reactions (3)-(10) [34, 36, 39]. Initial concentrations of PMS or lindane may have influences on the efficiency of UV/PMS process by way of influencing those reactions. As shown in Figure 2, the removal efficiency of lindane was found to increase significantly with increasing the initial concentrations of PMS, with a percentage removal of 58, 92, 98 and 100% in 180 min obtained, respectively, at an initial PMS concentration of 125, 250, 500 and 1000 µM. This trend can be explained by reaction (1), with the rate of radical formation increasing with an increase in the concentration of PMS [39]. An eight-time increase in PMS concentration led to a 6.2-time rise in the value of kobs, i.e., 4.77 × 10−3 min−1 at 125 µM PMS vs 2.97 × 10−2 min−1 at 1000 µM PMS (inset of Figure 2), demonstrating an inhibition role of a higher concentration of PMS. Possible explanations for this observation include the interaction of the reactive radicals with PMS following reactions (7) and (8), and the self-combination reactions among the reactive radicals (reactions (3)-(5)) especially at higher initial concentrations of the oxidant [39]. • + • → ( = 4 × 10 )
[40]
(3)
•
+ • → ( = (0.95 ± 0.08) × 10 )
[41]
(4)
•
+ • → ( = 5.3 × 10" )
[42]
(5)
• + → # + + •
[36]
(6)
• + → • + + # ( < 1.0 × 10 )
[43]
(7)
[36]
(8)
•
+ → • +
8
• + → + • ( = 8.3 × 10% , ' > 11) •
+ → • + ( = 1.2 × 10 , *+*+,-. ')
[36]
(9)
[34]
(10)
As seen in the Figure 3 and Table 1, with an increasing lindane concentration from 0.68 to 3.43 and 6.86 µM, the percentage degradation of lindane decreased from 100% in 120 min to 92 and 80%, respectively, in 180 min; while the kobs decreased from 3.51 × 10−2 to 1.42 × 10−2 and 7.89 × 10−3 min−1, respectively. The corresponding half-life values of these reactions were calculated and found to be 19.7, 48.8, and 87.8 min. A plausible reason could be from the formation of a large number of by-products (as will be discussed below on Figure 5) which competed with lindane for the reactive radicals and the concentration of which increased with increasing the initial concentration of lindane [44]. 3.3 Influence of water quality parameters (i.e., pH, humic acid or inorganic anions) Reactive radicals can react with both organic and inorganic compounds. Common water quality parameters, such as pH, inorganic anions and humic acids, may therefore compete with target compound for radical species leading to positive or negative impacts on the degradation kinetics [45]. Therefore, it is important to consider and evaluate the impact of these variables on the removal efficiency of lindane by UV/PMS. The SO4•− reactions such as those shown in reactions (6) and (7) could lead to the generation of hydronium ions (H3O+) which could subsequently result in a decrease in solution pH, as measured by the pH of 3.8, 5.4 and 7.2 for the initial solution pH of 4.0, 5.8 and 8.0, respectively, after 180 min. As shown in Figure 4, the removal efficiency was affected by varying the initial solution pH, with 86, 92 and 55% removal of lindane at initial pH 4.0, 5.8 and 8.0, respectively, in 180 min. The slight difference in the removal efficiency (~6%) by changing the pH from 4.0 to 5.8 was probably due to the unchanged radical quantum yield during PMS 9
photolysis in this pH range [27]. The decreased removal efficiency at higher solution pH, on the other hand, could be attributed to the scavenging of SO4•− and •OH by HO− in accordance with reactions (9) and (10), respectively [34, 36]. The effect of humic acid on the degradation of lindane by UV/PMS process is also illustrated in Figure 4. The addition of 1.0 mg L−1 humic acid decreased greatly the removal efficiency of UV/PMS, i.e., from 92 to 61% after 180 min, which could result from the scavenging effect of NOM towards •OH and SO4•− via reactions (11) and (12)), respectively [46, 47]. This result is consistent with the literature reports showing that efficiency of the UV/PMS system for atrazine removal decreased from 98% to 23% when 3.2 mg L−1 of humic acid was added to the reaction mixture [48]. Therefore, when applying UV/PMS AOPs, a pre-treatment aiming to remove organic matter may be considered in order to minimize the negative effects from those non-target compounds. The effect of inorganic anions on the efficiency of UV/PMS process was also investigated. The anions selected included CO32−, HCO3−, SO42− and Cl−, the presence of which led to different degrees of inhibition on the percentage removal of lindane, as shown in Figure 4. The presence of HCO3− or CO32− caused a significant inhibitory effect, attributable to the relatively strong scavenging of reactive radicals by these anions (reactions (13) – (16)) [34, 39]. The higher reactivity of CO32− with SO4•− and •OH than HCO3− allowed the former to show a higher inhibitory effect on the removal of lindane. Cl− can have a dual role on the efficiency of photochemical degradation processes, attributed to the transformation of SO4 •− and •OH into reactive but strongly selective radicals, e.g., Cl•, Cl2•− and ClOH•−, inducing either a promoting or an inhibiting effect on the removal of the organics by activated persulfate or peroxymonosulfate [49, 50]. Some of the important aqueous phase reactions of Cl• and Cl2 •− are shown in reactions 10
(17)-(27) [51, 52]. Thus despite the high reactivity of Cl− towards •OH and SO4•− (reactions (17) and (18)) [49], a negligibly small adverse effect was generally observed in this study on the efficiency of UV/PMS process. The overall removal efficiency of the UV/PMS process at an initial pH of 5.8 decreased in the following order: control ~ SO42− ~ Cl− > humic acid ~ HCO3− > CO32−. •
+ / → 0123456 ( = 2.23 × 10 7 (82+ 9 ) )
[46]
(11)
• + / → 0123456 ( > 6 × 10; 7 (82+ 9) )
[47]
(12)
9< + • → 9< • + ( = 4.1 × 10; )
[53]
(13)
9< + • → 9< • + + # ( = 3.5 × 10; )
[54]
(14)
9< + • → + 9< • ( = 3.9 × 10 )
[34]
(15)
9< + • → + 9< • ( = 8.5 × 10; )
[34]
(16)
9+ + • → 9+ • + ( = 2.6 × 10 )
[53]
(17)
9+ + • → 9+• ( = 4.3 × 10" )
[34]
(18)
9+• → 9+ + • ( = 6.1 × 10" )
[51]
(19)
9+ • + 9+ → 9+ • ( = 8.5 × 10" )
[52]
(20)
9+ • + 9+ • → 9+
[52]
(21)
9+ • + 9+ • → 9+ + 9+
[52]
(22)
9+ • + 9+ • → 9+ + 29+
[52]
(23)
9+ • + → 9+ • (2.5 × 10 )
[52]
(24)
9+ • + → 9+ • + 9+ (1.3 × 10< )
[52]
(25)
9+ • → 9+ • + 9+ (6.0 × 10 )
[51]
(26)
9+ • + → 9+ • + 9+ (4.0 × 10; )
[51]
(27) 11
3.4 By-product analysis and potential reaction mechanism With the kinetics being demonstrated to be feasible, the mechanism of the degradation of lindane by UV/PMS was assessed and presented in this section. There were mainly six reaction by-products identified by GC-MS in this study: 1,1,2,3,4,5,6-heptachlorocyclohexane (HeCH), 1,2,3,4,5,6-hexachlorobenzene (HCB), 1,3,4,5,6-pentachlorocyclohexene (PCCH), 3,4,5,6tetrachlorocyclohexene (TeCCH), 1,2,4-trichlorobenzene (TCB) and 2,4,5-trichlorophenol (TCP). These by-products have been reported previously in various oxidative studies on lindane, e.g., HeCH, HCB, PCCH, TeCCH and TCP in POMs photocatalysis [13], TCB and HCB in the photo-Fenton reaction [15], as well as HeCH, PCCH and TeCCH in TiO2 photocatalysis [55]. In this study, the exact reacting radical could not be distinguished with certainty because of the coexistence of both •OH and SO4•− in the reaction solution [31, 56]. Due to the similarities in the reaction mechanism of these two radicals, it is very likely that the detected by-products can be from either radical reaction. A potential reaction pathway for the degradation of lindane was proposed and is shown in Figure 5, including dechlorination, chlorination, dehydrogenation and hydroxylation. Dechlorination was presumably resulted from homolytic scission of the C-Cl bond upon UV excitation [57, 58]. As a result, chlorine radical (Cl•) was released from lindane, leaving a carbon centered radical. A subsequent abstraction of hydrogen from an adjoining carbon atom by •
OH and/or SO4•− might lead to the formation of PCCH. A further loss of chlorine resulted in the
formation of lesser chlorinated by-products such as TeCCH and TCB. Chlorination of lindane could result from its reaction with chlorine radical (Cl•) or dichlorine radical anion (Cl2•−). Cl• and Cl2•− are strong oxidizing species with E0 of 2.4 and 2.0 V, respectively [51]. These oxidizing species may react with organic compounds via addition to 12
double bond, hydrogen abstraction or electron-transfer reactions [59]. Thus an abstraction of hydrogen from lindane via Cl• led to the formation of HeCH. A similar explanation has been provided by Antonaraki et al. [13] employing POMs photocatalysis of lindane. In fact, the interaction of Cl− with SO4•− with the subsequent formation of the chlorinated organic compounds in water has been commonly reported in literature [60]. It has been widely reported that chlorine group was mainly responsible for the toxicity of chlorinated organic compounds [61]. The dechlorination achieved in water treatment processes may therefore result in a reduction in the overall toxicity [61]. However, with the chlorination to be one of the most significant reaction pathways in the transformation of lindane in UV/PMS, whether the toxicity of the reaction solution was reduced needs more investigations. Dehydrogenation might occur with the abstraction of two adjoining hydrogen atoms by •
OH and/or SO4•− attack. The resulting conjugated double bonded HCB by-product with more
thermal stability allowed the dehydrogenation to be kinetically feasible [62]. The formation of HCB from lindane via hydrogen abstraction by •OH was also reported elsewhere [13]. Hydroxylation is a process that introduces a hydroxyl group (-OH) into an organic compound. The addition of the electrophilic •OH to TCB formed a carbon centered radical, which by addition of O2 yielded a peroxy radical. After releasing HO2 •, TCP could be formed (reaction (28)). Hydroxylation of chlorobenzene and other chlorinated aromatic compound through such a pathway has been proposed earlier [63, 64].
(28)
13
In SO4•− mediated mechanisms, SO4•− oxidized the aromatic ring to a radical cation, which upon hydrolysis led to the formation of hydroxycyclohexadienyl radical. The resulting radical, after reaction with O2 and releasing subsequently HO2•, was converted into a hydroxylated phenolic by-product (reaction (29)) [30, 60, 65]. It should be noted, however, there were different positions available at the ring for hydroxylation reaction, which could be influenced by both steric hindrance and resonance stability [56, 66]. More studies are needed to differentiate the exact structure of TCP.
(29) Though not identified by our method, ring opening and cleavage by-products were also expected to form. The intermediate by-products HCB and TCP, for example, were known to mineralize into CO2, H2O and Cl− with an extended reaction time by photocatalytic and photochemical transformations [67, 68]. 3.5 Mineralization study As shown in the inset of Figure 6, though much less effective than the removal of the parent compound, a 56% TOC removal (TOC0 = 1.24 mg/L) by UV/PMS in 360 min could still be obtained. Increasing initial PMS concentration had an enhancing effect on lindane mineralization, with a 36, 49, 56 and 60% TOC removal achieved at an initial PMS concentration of 125, 250, 500, and 1000 µM, respectively, in 360 min (Figure 6). Ahmed and Chiron [33] also observed a similar mineralization trend during the oxidation of ciprofloxacin by UV/PMS. This could again be attributed to the increased concentration of •OH and/or SO4•− with 14
an increasing concentration of PMS, consistent with the effect of PMS concentrations on the degradation of lindane. 4 Conclusions In this study, the degradation of the environmentally persistent chlorinated pesticide lindane was investigated both kinetically and mechanistically. Compared to common hydroxyl radical reaction, SO4•− exhibited a slightly higher reactivity towards lindane as determined from its second-order rate constant of 1.3 × 109 M−1 s−1. Process parameters such as initial concentrations of PMS or lindane, as well as common water quality parameters such as organic humic acid or inorganic anions, showed a varied degree of impacts on the observed pseudo firstorder rate constant (kobs). The presence of humic acid, CO32− or HCO3− caused a strong inhibiting effect while the presence of SO42− or Cl− exerted a negligible effect on the efficiency of UV/PMS process. Various degradation by-products as identified by GC-MS revealed dechlorination, chlorination, dehydrogenation and hydroxylation to be potential transformation steps. Ring opening and cleavage could also be achieved as demonstrated indirectly by the significant decrease in TOC. This study shows UV/PMS based AOPs to be an effective method for the removal of lindane from aqueous environment. Acknowledgments The Higher Education Commission (HEC), Islamabad, Pakistan is highly acknowledged for funding this research project through an International Research Support Initiative Program (IRSIP). This work was also partially funded by the Cyprus Research Promotion Foundation through Desmi 2009-2010 which is co-funded by the Republic of Cyprus and the European Regional
Development
Fund
of
the
EU
under
contract
number NEA
IPODOMI/STRATH/0308/09. 15
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Microcystin-LR, Environmental Science & Technology, 42 (2008) 8877-8883. [67] A. Hiskia, E. Androulaki, A. Mylonas, S. Boyatzis, D. Dimoticali, C. Minero, E. Pelizzetti, E. Papaconstantinou, Photocatalytic Mineralization of Chlorinated Organic Pollutants in Water By Polyoxometallates. Determination of Intermediates and Final Degradation Products, Research on Chemical Intermediates, 26 (2000) 235-251. [68] Y.R. Wang, W. Chu, Photo-assisted degradation of 2,4,5-trichlorophenol by ElectroFe(II)/Oxone® process using a sacrificial iron anode: Performance optimization and reaction mechanism, Chemical Engineering Journal, 215-216 (2013) 643-650.
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Figure Captions Figure 1. Degradation of lindane by UV/PMS. [Lindane]0 = 3.43 µM, [PMS]0 = 250 µM, pH = 5.8. Figure 2. Effect of initial concentrations of PMS on the efficiency of UV/PMS process. Inset: variation of kobs (min−1) with [PMS]0. [Lindane]0 = 3.43 µM, pH = 5.8. Figure 3. Effect of initial concentrations of lindane on the efficiency of UV/PMS process. [Lindane]0 = 0.68, 3.43, and 6.86 µM, [PMS]0 = 250 µM, pH = 5.8. Figure 4. Effect of initial pH, humic acid or inorganic anions on the percentage removal of lindane in 180 min. [Lindane]0 = 3.43 µM, [PMS]0 = 250 µM, [humic acid]0 = 1.0 mg L−1, [inorganic anions]0 = 1000 µM, pH = 5.8 except noted. Figure 5. Proposed degradation pathway of lindane by UV/PMS. [Lindane]0 = 17.15 µM, [PMS]0 = 500 µM, pH = 5.8. Figure 6. Plots of TOC/TOC0 vs time by UV/PMS at using different initial concentrations of PMS, i.e., [PMS]0 = 125-1000µM, TOC0 = 1.24 mg/L. Inset: TOC vs lindane degradation : [Lindane]0 = 17.15 µM, [PMS]0 = 500 µM, pH = 5.8.
Table Captions Table 1. Pseudo first-order rate constant (kobs) and half-life (t1/2) at different initial concentrations of lindane in UV/PMS process.
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Figure 1.
1.0
C/C0
0.8
PMS only UV photolysis UV/PMS
0.6
0.4
0.2
0.0 0
50
100
150
200
Time (min)
26
Figure 2.
[PMS]0 = 125 µM 1.0
[PMS]0 = 500 µM [PMS]0 = 1000 µM
0.8
kobs (min-1)
0.03
[PMS]0 = 250 µM
0.02 0.01 0.00
C/C0
0
200 400 600 800 1000
0.6
[PMS]0 (µM)
0.4
0.2
0.0 0
50
100
150
200
Time (min)
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Figure 3.
[Lindane]0 = 6.86 mM [Lindane]0 = 3.43 mM [Lindane]0 = 0.68 mM
Lindane concentration (µM)
8
6
4
2
0
0
50
100
150
200
Time (min)
28
Figure 4.
Removal efficiency (%)
100
80
60
40
20
0 Humic acid CO32-
-
HCO3
Cl-
SO42-
Control pH = 4.0 pH = 8.0 (pH = 5.8)
Effect of water quality parameters
29
Figure 5.
30
Removal efficiency
Figure 6.
TOC/TOC0
1.0
0.8
1.0
TOC Lindane
0.8 0.6 0.4 0.2 0.0 0
100
200
300
400
Time (min)
0.6 125 µM 250 µM 500 µM 1000 µM
0.4
0.2 0
100
200
300
400
Time (min)
31
Table 1. ______________________________________________________________________________ [lindane]0 (µM)
Rate constant (min−1)
t1/2 (min)
R2
0.68
3.51 × 10−2
19.7
0.992
3.43
1.42 × 10−2
48.8
0.987
6.68
7.89 × 10−3
87.8
0.981
32
Graphical abstract
Lindane removal efficiency (%)
100
80
60
40
20
0
M NO
m (1
) g/L
2-
3 CO
(
M) 1m
- ( O3 C H
M) 1m
- (1 Cl
) mM
2-
4 SO
) ) ) Cl OH M) 5.8 H a ( m H (N .0 (p (1 .0 ol =4 8 r t = n pH Co pH
Effecet of water quality parameters
33
Highlights
•
Removal of lindane by UV-C/peroxymonosulfate based AOPs was investigated.
•
Lindane showed a comparable or slightly higher reactivity toward SO4•− than HO•.
•
Degradation efficiency was significantly affected in presence of NOM or alkalinity.
•
A reasonable mineralization of lindane was achieved by UV/PMS.
•
Transformation mechanism was proposed based on by-products identified via GC-MS.
34