Photodegradation kinetics of iopamidol by UV irradiation and enhanced formation of iodinated disinfection by-products in sequential oxidation processes

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Photodegradation kinetics of iopamidol by UV irradiation and enhanced formation of iodinated disinfection by-products in sequential oxidation processes Fu-Xiang Tian a, Bin Xu a,*, Yi-Li Lin b, Chen-Yan Hu c, Tian-Yang Zhang a, Nai-Yun Gao a a State Key Laboratory of Pollution Control and Resources Reuse, Key Laboratory of Yangtze Aquatic Environment, Ministry of Education, College of Environmental Science and Engineering, Tongji University, Shanghai 200092, PR China b Department of Safety, Health and Environmental Engineering, National Kaohsiung First University of Science and Technology, Kaohsiung 824, Taiwan, ROC c College of Environmental and Chemical Engineering, Shanghai University of Electric Power, Shanghai 200090, PR China

article info

abstract

Article history:

The photochemical degradation of iopamidol with low-pressure UV lamps and the for-

Received 7 January 2014

mation of iodinated disinfection by-products (I-DBPs) during sequential oxidation pro-

Received in revised form

cesses including chlorine, monochloramine and chlorine dioxide were investigated in this

26 March 2014

study. Iopamidol can be effectively decomposed by UV irradiation with pseudo-first order

Accepted 27 March 2014

reaction kinetics. The evaluated quantum yield was found to be 0.03318 mol einstein
1.

Available online 5 April 2014

Results showed that iopamidol degradation rate was significantly increased by higher UV intensity and lower initial iopamidol concentration. However, the effect of solution pH was

Keywords:

negligible. Degradation of iopamidol by UV photolysis was subjected to deiodination and

Iodinated disinfection by-products

hydroxylation mechanisms. The main degradation products including eOH substitutes

(I-DBPs)

and iodide were identified by UPLC-ESI-MS and UPLC-UV, respectively. Increasing the in-

UV irradiation

tensity of UV irradiation promoted the release of iodide. Destruction pathways of iopa-

Chlorine

midol photolysis were proposed. Enhanced formation of I-DBPs were observed after

Monochloramine

iopamidol photolysis followed by disinfection processes including chlorine, monochlor-

Chlorine dioxide

amine and chlorine dioxide. With the increase of UV fluence, I-DBPs formation were

Kinetics

significantly promoted. ª 2014 Elsevier Ltd. All rights reserved.

* Corresponding author. Tel.: þ86 13918493316. E-mail addresses: [email protected], [email protected] (B. Xu). http://dx.doi.org/10.1016/j.watres.2014.03.069 0043-1354/ª 2014 Elsevier Ltd. All rights reserved.

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1.

Introduction

Recently, iodinated disinfection by-products (I-DBPs), which mainly refer to iodo-acids (I-acids) and iodo-trihalomethanes (I-THMs), have drawn great concerns as they are predicted to be much more toxic than classical DBPs (Bichsel and von Gunten, 2000; Richardson and Postigo, 2012; Smith et al., 2010; Woo et al., 2002). I-DBPs highly enhanced mammalian cell cytotoxicity and genotoxicity as compared to their brominated and chlorinated analogues (Richardson et al., 2008). For example, iodoacetic acid, one of the five I-acids identified, is the most genotoxic DBPs studied to-date in mammalian cells (Plewa et al., 2004; Richardson and Postigo, 2012). Besides, I-THMs (especially iodoform) are responsible for bad taste and odor in drinking waters (Bichsel and von Gunten, 2000). An occurrence study conducted in the United States and Canada showed that I-acids and I-THMs were found in effluents from most plants at maximum levels of 1.7 mg L
1 (iodoacetic acid), 1.4 mg L
1 (bromoiodoacetic acid), 10.2 mg L
1 (bromochloroiodomethane), 7.9 mg L
1 (dichloroiodomethane) (Richardson et al., 2008). A research indicated that I-THMs (especially iodoform) were the main products in chloramination of iodide-containing waters (Bichsel and von Gunten, 2000). Iodinated X-ray contrast media (ICM) are widely used in human medicine for imaging of organs or blood vessels during diagnostic tests (Duirk et al., 2011; Perez et al., 2006). As worldwide consumption reaching 3.5  106 kg year
1 (Perez and Barcelo, 2007), ICM have been frequently detected in effluents of wastewater treatment plants (WWTPs) and surface waters at elevated concentrations due to their incomplete removal (Perez et al., 2006; Schulz et al., 2008; Sugihara et al., 2013), with 35e55% of non-ionic ICMs and below 20% of ionic ICM removal by ozonation in drinking water treatment plants (DWTPs) (Seitz et al., 2008). Although ICM are not toxic to human bodies, their presence in source waters may lead to the formation of highly toxic I-DBPs in chlorinated and chloraminated drinking water (Duirk et al., 2011). Besides, the high adsorbable organic iodine (AOI) in municipal treatment plant effluents and surface waters are identified to be derived from ICM (Putschew et al., 2000). Advanced treatment processes have shown some success in degrading ICM, attributed to reactions with hydroxyl radicals (Seitz et al., 2008; Ternes et al., 2003). Studies on the treatment of ICM by advanced oxidation processes (AOPs) such as g-irradiation, UV/H2O2, UV/TiO2 and O3/H2O2 have been reported (Doll and Frimmel, 2004; Huber et al., 2005; Jeong et al., 2010; Sugihara et al., 2013; Ternes et al., 2003) and the rate constants for reactions of hydroxyl radical with some ICM were determined (Jeong et al., 2010). Iopamidol is the most frequently detected ICM in aqueous environment (Duirk et al., 2011; Perez et al., 2006) with concentration up to 2.7 mg L
1 in a raw water in USA (Duirk et al., 2011), 15 mg L
1 in effluents of WWTPs, 2.4 mg L
1 in groundwater and 0.49 mg L
1 in rivers in Germany (Ternes and Hirsch, 2000). Until now, iopamidol has been supposed to be the most important contributor of ICM to I-DBPs formation (Duirk et al., 2011). Some researchers have proposed the aquatic formation mechanisms of I-DBPs. The source of iodine in I-DBPs is

199

believed to be from iodine-based disinfectants (Smith et al., 2010), naturally occurring iodide (Bichsel and von Gunten, 1999, 2000; Ye et al., 2013, 2012) and iodinated organic compounds such as ICM consumed (Duirk et al., 2011). Smith et al. (2010) proved that iodoform was the predominant THMs formed during iodination and it got a higher yield during the treatment with iodine tincture than iodine tablets. Many researchers have studied the fates of iodide during treatment by different oxidants such as chloramine, chlorine, ozone, potassium permanganate and chlorine dioxide (Allard et al., 2013; Bichsel and von Gunten, 1999; Fabian and Gordon, 1997; Kirschenbaum and Sutter, 1966; Lengyel et al., 1993, 1996). It was reported that during oxidative treatment of iodide-containing waters with chlorine or chloramine, hypoiodous acid (HOI) was quickly formed and then reacted with natural organic matter (NOM) to produce I-THMs (Bichsel and von Gunten, 2000). Allard et al. (2013) demonstrated that ozone pre-treatment selectively oxidized iodide to iodate and avoided the formation of I-DBPs. Owing to its immediate and complete reduction to iodide in vivo, iodate formed during oxidation is non-toxic (Taurog et al., 1966). Tao et al. believed that when iodide-containing waters were treated with chlorine dioxide, I
is oxidized to I2. Then the presence of active iodine species (HOI, I2 and I
3 ) might react with NOM, resulting in the formation of I-DBPs (Ye et al., 2013). Another study by Ye et al. (2012) indicated that treatment of iodide-containing water with potassium permanganate could also lead to the formation of I-DBPs. Duirk et al. (2011) evidenced that ICM can constitute an organic iodine source to form I-DBPs in chlorinated and chloraminated especially NOM-containing source waters. Due to the disinfection effectiveness against a wide range of waterborne pathogens, producing no regulated DBPs and the continuous decrease of treatments costs, UV is becoming more and more widespread in wastewater and drinking water disinfection (Kashinkunti, 2004; Pereira et al., 2007). Sequential disinfection using UV and chlorination are expected to minimize DBPs formation because lower chlorine dosages can be used (Liu et al., 2012). One major problem encountered with UV disinfection is that it cannot provide any residual disinfection once the effluents have left the UV photoreactor (Hassen et al., 2000). Therefore, application of chlorine-based disinfectant (chlorine, monochloramine and chlorine dioxide) so as to resupply residual disinfection after UV is technically necessary in water treatment process. However, a few literature have reported that UV irradiation may enhance the formation of dichloroacetic acid, trichloroacetic acid, cyanogen chloride (Liu et al., 2006, 2012) and chloropicrin (Shah et al., 2011), or induce transformation of halobenzoquinones (Qian et al., 2013) and N-nitrosodimethylamine (Lee et al., 2007; Soltermann et al., 2013). Nevertheless, there are no reports relating to effects of UV treatment to the formation of IDBPs and little research has been conducted about UV degradation removal of ICM. The objective of this study is to (1) investigate the degradation kinetics and pathways of iopamidol during UV irradiation, and (2) to evaluate the effects of UV treatment on the formation of classical- and I-DBPs during subsequent oxidation processes by chlorine, monochloramine and chlorine dioxide. This study highlighted the possible mechanism of ICM

200

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photodegradation and how these contaminants transformed to toxic I-DBPs in waters.

2.

Materials and methods

2.1.

Chemicals

All chemicals were at least of analytical grade except as noted. The calibration standards, internal standards, surrogate standards for volatile DBPs (including THMs with 3 hydrogen atoms being replaced by chlorine or bromine (THM4), vinyl chloride (VC), haloketones (HKs), haloacetonitriles (HANs) and chloropicrin (CP)), EPA 552.2 haloacetic acids mix, CHI3 (99%), iodoacetic acid (IAA) (99.0%) standards, sodium hypochlorite (NaOCl) solution (available chlorine 4.00e4.99%), ammonium acetate (98%), NaOH (98%), KH2PO4 (99.0%), Na2CO3 (99.0%), NaHCO3 (99.0%) and KI(99.0%) were purchased from SigmaeAldrich (St. Louis, MO, USA). The CHClI2, CHBrI2, CHCl2I, CHBr2I and CHBrClI standards were obtained from CanSyn Chemical Corp. (Canada). Iopamidol (99.6%) was obtained from U.S.Pharmacopeia. Formic acid solution (49e51%) was purchased from Fluka (St. Louis, MO, USA). triiodoacetic acid (TIAA) (90%) standard solution was obtained from Toronto Research Chemicals Inc.(Canada). Methyl tert-butyl ether (MtBE) and acetonitrile were obtained from J.T.Baker (USA). Atrazine was obtained from Dr. Ehrenstorfer (German) in chromatographical purity (>99.0%) and used without further purification. Analytical grade reagents including Na2S2O3, H2O2, K2TiO(C2O4)2 and H2SO4 were purchased from Sinopharm Chemical Reagent Co., Ltd. (Shanghai, China) without further purification. All solutions were prepared with ultrapure water produced from a Milli-Q water purification system (Millipore, USA). A monochloramine (NH2Cl) solution was freshly prepared by mixing aqueous ammonium chloride and NaOCl solutions at a Cl2/N molar ratio of 0.8 at pH 8.5. The 0.3% of ClO2 stock solution was prepared as detailed elsewhere in our previous study (Ye et al., 2013).

2.2.

UV irradiation

The UV irradiation system contained a 9 L stainless steel reactor (i.d. ¼ 20 cm, height ¼ 32 cm) with a 400 mL quartz tube (i.d. ¼ 4.5 cm) fixed in the center. The temperature of the UV system was controlled at 25  C by connecting the system to a low temperature water bath (DKB-1915, Jinghong, Shanghai, China) and circulating the water by a diaphragm pump (Model DP0700, Puricom, Taiwan, China). Four low pressure (LP) Hg UV lamps (TUV 11 W T5 4P-SE, Philips, Netherlands) with quartz sleeves were symmetrically fixed in the center of the reactor. The UV intensity was changed by turning on different numbers of UV lamps and detected using a light intensity meter (UVeC luxometer, Photoelectric Instrument Factory of Beijing Normal University, Beijing, China) at 254 nm. The UV fluence (mJ cm
2) delivered to the sample was calculated as the UV intensity multiplied by the exposure time (Yuan et al., 2009; Rosenfeldt et al., 2006). UV fluence rate values were determined using atrazine actinometer (Canonica et al., 2008) with a quantum yield of

0.046 mol einstein
1 and molar extinction coefficient of 3860 M
1 cm
1 at a wavelength of 254 nm. According to Equation (1) (Beltran et al., 1995), when the optical density (2.3LεCt) is greater than 2, a plot of C0
Ct versus time should lead to a straight line of slope k ¼ 2.3I0F. Fig. S1 of Supplementary Material presented such a plot obtained from results of actinometer experiments. It can be seen that the curves all showed a good linearity (R2 > 0.99) with the present photoreactor setup. Then the fluence rate values corresponds to the UV intensity of 1.53, 3.02, 4.45 and 5.96 mW cm
2, were determined as 6.891  10
7, 14.159  10
7, 18.879  10
7, 25.487  10
7 E s
1 L
1, respectively. C0
Ct ¼ 2:3I0 Ft

(1)

C0 ¼ 2:3LI0 Fεt Ct

(2)

ln

where C0 and Ct are the concentrations of the organic material at the reaction time of t ¼ 0 and t ¼ t (M); I0 the UV fluence rate value (einstein s
1 L
1); F the quantum yield of photodegradation for UV (mol einstein
1); ε the molar extinction coefficient of organic material for UV (M
1 cm
1); L the optical path length (cm). The path length of the reactor were determined by hydrogen peroxide actinometry with a quantum yield of 1.0 mol einstein
1 and molar extinction coefficient of 19 M
1 cm
1 at a wavelength of 254 nm by Equation (2) (Beltran et al., 1995), which indicates a first-order kinetics for the actinometer (when the optical density (2.3LεCt) is lower than 0.2). Hence a plot of lnC0/Ct versus time has to be linear and the slope was k ¼ 2.3LI0Fε. Fig. S2 of Supplementary Material showed this plot corresponding to two experiments conducted at low initial hydrogen peroxide concentrations. Thus, with the UV fluence rate values obtained, the effective path length of the reactor with 1, 2, 3 and 4 lamps open were found to be 7.225, 6.405, 6.616 and 6.443 cm, respectively.

2.3.

Experimental procedures

The iopamidol solutions were beforehand buffered using 10 mM phosphate (for pH 5e8) or 10 mM carbonate (for pH 9) buffer. pH values were adjusted with small volume of 0.01, 0.1 or 1 M H2SO4 and NaOH. In UV degradation experiment, a 200mL solution spiked with iopamidol at an established concentration and buffered at a certain pH with temperature controlled at 25  1  C was subjected to UV irradiation. 1 mL of solution was rapidly transferred into a UPLC vial at different time-points, and then the sample was analyzed by ultra performance liquid chromatography (UPLC) or UPLC-ESI-MS immediately. Because our preliminary experimental results showed almost completely removal of iopamidol with initial concentration of 10 mM at pH 7 within 5 min UV exposure, the exposure time was set accordingly. After exposing to certain UV intensity (ranging from 0 to 1788 mJ cm
2) for 5 min at pH 7, iopamidol solution was dosed with 100 mM chlorine, monochloramine or chlorine dioxide. Solution pH was adjusted to 7.0 using NaOH and H2SO4 solution. Then DBPs formation experiments were conducted in triplicate under headspacefree conditions in 40-mL glass screw-cap amber vials with

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PTFE-lined septa in dark at temperature (25  1  C) in a thermostatic biochemical incubator. At the scheduled time of 7 d, the reaction was quenched using NH4Cl (for Cl2 and NH2Cl) or Na2S2O3 (for ClO2) solutions at 1.2 times of oxidants concentration and the samples were then extracted for the analysis classical DBPs and I-DBPs. The analysis of iopamidol degradation intermediates experiments were conducted during UV irradiation process at a concentration of 25 mM in order to enhance the production of degradation intermediates for higher analytical resolution. At the scheduled time, samples were collected and analyzed as soon as possible.

2.4.

201

of an Accela U-HPLC system and a TSQ Quantum mass spectrometer (ESI source, Thermo Scientific Inc.,USA). An XTerraR MS C18 column (250 mm  2.1 mm, i.d., 5 mm film thickness, Waters, USA) was used for product analysis. Acetonitrile (10%) and 1% formic acid solution (90%) were used as the mobile phase with a flow-rate of 0.25 mL min
1. Operating parameters of the ESI conditions were as follows: spray voltage 4.0 kV, capillary temperature 350  C, sheath gas pressure 50 (arbitrary units), and auxiliary gas pressure 45 (arbitrary units). MS chromatograms were obtained both in total ion current (TIC) mode, using full scans (m/z 200e900) for mass spectra acquisition and in the selected ion monitoring (SIM) mode.

Analytical methods

Iopamidol, atrazine and iodide released from photolysis of iopamidol (Hurst et al., 1983) were all analyzed using UPLC (Waters, USA) equipped with a XTerra MS C18 column (4.6  250 mm i.d., 5 mm film thickness, Waters, USA) by an UV detector at wavelength of 242 nm, 223 nm and 226 nm, respectively. The mobile phase for iopamidol was consisted of 10%/90% (v/v) acetonitrile and Milli-Q water at a flow rate of 0.80 mL min
1. The mobile phase for atrazine was consisted of 60%/40% (v/v) acetonitrile and Milli-Q water at a flow rate of 0.80 mL min
1. The mobile phase for iodide was consisted of 2%/98% (v/v) acetonitrile and KH2PO4 solution (0.09 M) at a flow rate of 0.80 mL min
1. The injection volume was 10 mL. The detection limit of iopamidol, atrazine and iodide were 5.0, 10.0 and 5.0 mg L
1, respectively. The pH measurements were carried out with a regularly calibrated pH meter (FE20-FiveEasy, Mettler Toledo, Switzerland) using standard buffer solutions (Mettler Toledo; pH ¼ 4.01, 7.00, 9.21). The concentration of hydrogen peroxide was determined spectrophotometrically by using the titanium oxalate method (De Laat and Gallard, 1999). The absorbances were measured using a 1-cm quartz cell at wavelength of 254 nm for iopamidol and 400 nm for the titanium peroxo complex (SQ-4802 UVeVis spectrophotometer, UNICO, Shanghai). Chlorine and monochloramine concentration were calibrated by the N, N-diethyl-p-phenylenediamine (DPD) colorimetric method (APHA, 1998). The concentration of ClO2 stock solution was measured with a DR/890 portable colorimeter (Hach, USA) after being diluted by Milli-Q water into an amber volumetric flask to avoid exposure to light. Program 112 with Mid Range (0.0e5.0 mg L
1 ClO2) was used and the estimated detection limit was 0.1 mg L
1 ClO2. The concentration of total organic carbon (TOC) was determined using TOC-L (Shimadzu, Japan) with ASI-L autosampler and samples were adjusted to weak acid conditions (pH 5e6) using 0.1M H2SO4 before analysis. The detection limit of TOC was 0.1 mg L
1. Methods for quantifying classical DBPs and I-DBPs were developed by modifying USEPA Method 551.1 (Munch and Hautman, 1995) and 552.2 (Hodgeson et al., 1995). Samples were extracted with MtBE and the extracts were analyzed using a gas chromatograph (GC-2010, Shimadzu, Japan) equipped with an electron capture detector (ECD) and an HP-5 capillary column (30 m  0.25 mm i.d., 0.25 mm film thickness, J&W, USA). Photodegradation intermediates of iopamidol during UV irradiation were identified by a UPLC-ESI-MS system consisted

3.

Result and discussion

3.1.

UV degradation kinetics

A number of authors have conducted studies on photooxidation of organic substances by UV irradiation and several kinetic models have been developed (Benitez et al., 2004; Kim et al., 2009; Lopez et al., 2003; Pereira et al., 2007) (Salgado et al., 2013; Yuan et al., 2009). Generally, the concentration decrease of an organic material by UV photodegradation versus time can be expressed as follows (Benitez et al., 2004; Lopez et al., 2003): dCt ¼ FI0 ½1
expð
2:3LεCt Þ
dt

(3)

The implications of these parameters were interpreted in Section 2.2. Integrating Equation (3), the following expression is obtained:   εLC0 10
1 ¼ 2:3FI0 Lεt ¼ kobs t ln εLC t 10
1

(4)

Here, kobs is the first order rate constant (s
1). Thus, according to Equation (4), a plot of the first term vs. reaction time must lead to a straight line, whose slope is a pseudo-first order rate constant kobs ¼ 2.3FI0Lε. The experimental results obtained in the iopamidol photooxidation experiments with different UV intensity were plotted in Fig. 1. Good linear fits (R2 > 0.98) of experimental data were indicative of pseudo-first order reaction kinetics of iopamidol degradation. After regression analysis, the slopes kobs also depicted in Fig. 1 were deduced. The results are consist with previous studies conducted by other researchers (Benitez et al., 2004; Kim et al., 2009; Lopez et al., 2003). It can also be seen from Fig. 1 that greater UV intensity leads to faster degradation of iopamidol in solution. Therefore, the effectiveness of UV irradiation for iopamidol elimination in water treatment process could be expected. As shown in Fig. 1, fine linear fits of kobs and UV fluence rate values were also observed (R2 ¼ 0.971), equation of which can be expressed as kobs ¼ 21617I0. Then, with the obtained value of ε (Pereira et al., 2007) (22,700 M
1 cm
1 for iopamidol) and those of I0, and L above provided in Section 2.2, the quantum yield (F) of iopamidol was evaluated as 0.03318 mol einstein
1. This value was a bit smaller than the quantum yield of iohexol, another ICM with a higher molar extinction coefficient of 27620 M
1 cm
1, which was reported to be

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Fig. 1 e Pseudo-first order kinetics plot of iopamidol degradation with UV intensity of 1.53, 3.02, 4.45, 5.96 mW cmL2, respectively. (pH [ 7, [phosphate buffer] [ 10 mM, [iopamidol]0 [ 10 m M.)

0.0403 mol einstein
1 by Pereira et al. (2007). The relatively high quantum yield of iopamidol observed is consistent with its high molar extinction coefficient and indicates a high photodegradation efficiency using direct photolysis (Salgado et al., 2013).

3.2. Effect of initial iopamidol concentration and solution pH The complicated Equation (4) can be simplified to Equation (1) and Equation (2) according to the value of the exponential term (Beltran et al., 1995; Lopez et al., 2003), i.e. the optical density (2.3LεCt), which depends on the concentration of the organic material in a certain reactor. The kinetics features of these two models are quite different. As the molar extinction coefficient of iopamidol is high, variations of initial concentration might affect the photodegradation rate of the contaminant. Hence, experiments on effect of different initial

concentrations of iopamidol ranging from 1.25 to 20 mM were conducted, results of which were presented in Table 1. As indicated in Table 1, the lower the initial concentration of iopamidol, the higher the degradation rate which is expected in homogeneous oxidation processes and this result is consistent with research by Lopez et al. (2003). It is deduced that lower concentrations of iopamidol in solution were expected to undergo faster degradation by UV irradiation. Experiments on the photodecomposition of iopamidol by UV irradiation were also carried out by modifying the pH (5, 6, 7, 8 and 9). The results obtained were also presented in Table 1. It is known that pH can affect the mechanism and pathways of degradation (Chiang et al., 2004). Besides, the distribution of reactants species, depending on their pKa values, may influence the light absorbance and photolytic properties under various solution conditions (Shen et al., 1995). However, as Table 1 depicted, no significant difference of kobs was observed among different pH values and it seemed that the degradation rate was not affected by the solution pH. This was probably due to the fact that at all investigated pHs, iopamidol is not dissociable in solution and there are no ionized species present, which indicated that the monochromatic absorption coefficient at 254 nm of the iopamidol solution will not change with pH values (Benitez et al., 2004). It is also known that pH has much effect on the oxidation capacity of chemical disinfectants such as chlorine, due to the dissociation of chlorine and varied reaction mechanism at different pH values (Xu et al., 2011). However, such things would not occur in UV irradiation systems if species of contaminant are stable forms (Benitez et al., 2004; Shen et al., 1995). This result also demonstrated that concentrations of hydrogen ion and hydroxyl ion were not expected to influence the reaction system of iopamidol and UV irradiation. Accordingly, changing acidity or basicity of waters will not improve the UV degradation rate of iopamidol.

3.3. Destruction pathways of iopamidol during UV irradiation It has been evidenced that the high-molecular-weight organic components of NOM become more aliphatic in character after

Table 1 e First order rate constants of iopamidol degradation during UV irradiation. Initial iopamidol concentration (mM) 10 10 10 10 10 10 10 10 10 20 10 5 2.5 1.25

UV intensity (mW cm
2)

pH

kobs (s
1)

t1/2 (s)

Iopamidol degradation after 5 min UV exposure (%)

R2

1.53 3.02 4.45 5.96 3.02 3.02 3.02 3.02 3.02 3.02 3.02 3.02 3.02 3.02

7 7 7 7 5 6 7 8 9 7 7 7 7 7

0.0124 0.0253 0.0434 0.0569 0.0233 0.0232 0.0245 0.0236 0.0231 0.0247 0.0247 0.0266 0.0317 0.034

163.6 72.9 43.6 32.5 79.2 79.5 75.3 78.2 79.9 136.9 74.7 44.9 29.1 23.6

81.4 98.3 100.0 100.0 99.4 99.4 99.5 99.5 99.4 98.1 98.2 99.7 100.0 100.0

0.985 0.994 0.992 0.982 0.980 0.981 0.989 0.980 0.977 0.993 0.977 0.986 0.988 0.988

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UV irradiation because the appearance of more carboxyl and carbonyl carbon atoms (Kulovaara et al., 1996). Previous study also revealed that the iodo group is a better leaving group than chloro group (Woo et al., 2002). Research on degradation of ICM by advanced oxidation/reduction processes (Jeong et al., 2010) and simulated solar radiation photolysis (Pe´rez et al., 2009) have provided evidence for the deiodination and hydroxylation mechanisms in transformation of these triiodinated compounds. Accordingly, deiodinated products are expected to be formed from UV irradiation. Moreover, intermediates with hydroxyl and carbonyl groups might also be present. Photodegradation intermediates of iopamidol by UV irradiation were analyzed by UPLC-ESI-MS with the total ion chromatograms (TIC) mode and the selected ion monitoring

203

(SIM) mode. Chromatograms obtained by UPLC-UV and UPLCESI-MS analysis were provided in Fig. S3 and Fig. S4 of Supplementary Material. Chromatographic retention times as well as mass spectral data for these proposed intermediates were listed in Table S1 of Supplementary Material. The molecular ion of iopamidol, corresponding to the (M þ H)/z peak, was 777. The intermediates of iopamidol short-written as M126, M-110, M-236, M-252, M-362, M-346 products, represented the net mass loss of the products from the parent compound. Postulated structures for the photodegradation products of iopamidol during UV irradiation were summarized in Fig. 2. The M-126 product was characterized as scission of one iodine atom and substitution of the halogen by a hydrogen atom. The stepwise deiodination of iopamidol was also accompanied by the addition of hydroxyl group at the iodo site on the aromatic

Fig. 2 e Proposed destruction pathways of iopamidol during UV irradiation.

204

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ring. So the ions of M-110, M-236, M-220, M-362, M-346 and M330 were then identified as different derivatives of iopamidol with partial or complete iodo sites substituted by hydroxyl groups. Besides, hydrogen atom abstraction reactions from side chains of deiodinatd intermediates forming ketone products have been proposed by Jeong et al. (2010). Pe´rez et al. (2009) also believed that the corresponding ketone was formed by oxidation of a methylene group in the hydroxylated side chain during phototransformation of iopromide. The detected ions of M-2, M-128, M-112, M-238, M-364, M-222 and M-332 also proved the formation of ketone products in degradation of iopamidol. These products identified above were in agreement with previous studies suggesting deiodination being the major route in decomposition of other triiodinated compounds such as iopromide, iohexol and diatrizoate (Jeong et al., 2010; Pe´rez et al., 2009; Sugihara et al., 2013). In summary, the destruction pathways of iopamidol during UV irradiation were proposed and illustrated in Fig. 2.

3.4. Deiodination and mineralization of iopamidol during UV irradiation From the analysis above, it was concluded that UV irradiation can induce gradual deiodination of iopamidol and iodide would be accordingly released. Limited reports demonstrated that irradiation of ICM resulted in the deiodination of the matrix and the release of I
in solution (Doll and Frimmel, 2004; Jeong et al., 2010; Steger-Hartmann et al., 2002). Simultaneous degradation and deiodination of iopamidol with UV intensity as 1.53, 3.02 and 4.45 mW cm
2, respectively, at initial iopamidol concentration of 10 mM, were shown in Fig. 3. As can be seen from Fig. 3, UV irradiation of iopamidol indeed resulted in the increasing release of I
as irradiation time went by, which indicated the stepwise deiodination of the triiodinated contrast media and were consistent with the identified intermediates using UPLC-ESI-MS. This result was also in agree with studies on degradation of other ICM by AOPs or photocatalytic process (Doll and Frimmel, 2004; Jeong et al., 2010; Pe´rez et al., 2009). The deiodination was more favorable

Fig. 3 e Simultaneous degradation and deiodination of iopamidol with UV intensity of 1.53, 3.02 and 4.45 mW cmL2, respectively. (pH [ 7, [phosphate buffer] [ 10 mM, [iopamidol]0 [ 10 mM.)

at higher UV intensity with 34.06, 41.57 and 55.75% of I
released from iopamidol at the UV intensity of 1.53, 3.02 and 4.45 mW cm
2, respectively. It was also shown in Fig. 3 that though iopamidol was completely degraded at 180 s, concentration of I
still increased as irradiation continued, which could be attributed to the stepwise deiodination of UV intermediates of iopamidol with iodo group in their structures. High percentages of I
released in solution from iopamidol and the intermediates formed after UV irradiation will enormously affect the quality and safety of water, in the sequential disinfection process, especially for the cases with higher UV intensity and longer irradiation time, which will be discussed in the following Section 3.5. During the UV irradiation of iopamidol, TOC was also monitored to better understand the degradation and mineralization of iopamidol during UV treatments. With applied UV intensity of 1.53, 3.02 and 4.45 mW cm
2, initial concentration of iopamidol as 10 mM, the time course of TOC during the photodegradation of iopamidol was shown in Fig.S5 of Supplementary Material, which indicated that concentrations of TOC did not significantly change during the 5-min UV irradiation. Although the concentration of iopamidol considerably decreased with time (Fig. 3), no oxidation products and intermediates transformed into CO2. Fractions of smaller molecular weight than the original molecule were supposed to be generated and the TOC concentration maintained almost constant throughout the irradiation time (Prados-Joya et al., 2011). This result also implied that UV can not mineralize iopamidol sample to the desired extent and the total organic carbon remained in solution will still lead to organic contamination and other potential threat to water safety, for example, providing carbon source for DBPs formation in sequential oxidation process.

3.5. Formation of classical DBPs and I-DBPs in UV irradiation and sequential oxidation processes As discussed in Section 3.2 and 3.4, UV irradiation could degrade iopamidol into many intermediates without minimerization to CO2. Deiodination and hydroxylation of the parent compounds might enhance the biodegradablity of ICM during water treatment processes (Jeong et al., 2010). However, what draws our concerns most is the fate of iodine derived from iopamidol deiodination. eOH substitutes of iopamidol during UV irradiation, which can provide organic carbon source for I-DBPs formation, might facilitate the subsequential attack by oxidants. As I
could be oxidized to HOI in the presence of oxidants, I-DBPs are expected to be formed by reactions of HOI with NOM (Bichsel and von Gunten, 2000). An additional research issue to be tackled is I-DBPs formation in oxidation followed by UV treatment. Classical DBPs and I-DBPs formation of reaction mixtures containing iopamidol solution pretreated by different UV irradiation intensities with Cl2, NH2Cl or ClO2 were exhibited in Fig. 4. The sequence numbers standing for different experimental conditions of UV pretreatment and oxidants were listed in Table 2 along with the summation of classicaland I-DBP formation and the conversion of iodine to I-DBPs for each set of experiment.

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Table 2 e Formation of classical- and I-DBPs from iopamidol during UV irradiation followed by different oxidation processes. Sequence numbera A1 B1 C1 D1 E1 A2 B2 C2 D2 E2 A3 B3 C3 D3 E3

Classical DBPsb formation (mM)

I-DBPs formation (mM)

Conversion of iodine to I-DBPs (%)

4.463 2.630 1.587 1.347 1.200 0.050 0.142 0.183 0.184 0.153 0 0.019 0.037 0.055 0.029

0.859 1.348 1.287 1.515 1.490 0.488 0.997 1.465 1.565 1.730 0 0.123 0.408 0.469 0.660

3.224 7.936 7.404 8.915 8.786 4.568 7.452 11.447 13.033 14.147 0 1.006 3.283 3.697 5.182

Notes: a Capital letters A, B, C, D and E indicate pretreatment by different UV fluence as 0, 459, 906, 1335 and 1788 mJ cm
2, respectively; subscript numbers 1, 2 and 3 represent subsequent oxidation by Cl2, NH2Cl and ClO2, respectively. b Classical-DBPs refer to the sum of CF, DCAN, TCAA and DCAA formation.

Fig. 4 e Formation of classical DBPs (a) and I-DBPs (b) during UV irradiation of iopamidol followed by different oxidation processes. (UV fluence of 0, 459, 906, 1335 and 1788 mJ cmL2, respectively; pH [ 7, [phosphate buffer] [ 10 mM, [iopamidol]0 [ 10 mM, [Cl2]0 [ [NH2Cl]0 [ [ClO2]0 [ 100 mM, t [ 7 d) Error bars represent one standard deviation of triplicate measurements.

As is shown in Fig. 4a the increase of UV fluence depress the formation of classical-DBPs (including chloroform (CF), dichloroacetonitrile (DCAN), trichloroacetic acid (TCAA) and dichloroacetic acid (DCAA) while significantly enhanced the formation of I-DBPs (Fig. 4b) during subsequential oxidation by Cl2, NH2Cl or ClO2. UV induced formation of CHI3 and TIAA (CI3COOH) to following disinfection processes. For the cases of UV irradiation followed by chlorination, notable decrease of total classical-DBPs from 602.0 mg L
1 (sequence A1 in Fig. 4a and Table 2) to 168.6 mg L
1 (sequence E1 in Fig. 4a and Table 2). DCAN and TCAA contribute to the majority of the measured classical-DBPs during chlorination of iopamidol both in the absence (sequence A1 in Fig. 4a and Table 2) and presence of UV irradiation (sequences B1, C1, D1 and E1 in Fig. 4a and Table 2). The results are relevant to the peptide bond groups in the molecular structure of iopamidol

that characterizes amino acids susceptible to form higher yields of HAAs and DCAN than THMs during chlorination (Hong et al., 2009; Trehy et al., 1986). However, limited amount of DCAN (5.552 mg L
1) was detected during monochloramination of iopamidol without UV irradiation (sequence A2 in Fig. 4a and Table 2). And UV irradiation also resulted in slight increase of DCAN formation (approximately 10e15 mg L
1) in sequential monochloramination(sequences B2, C2, D2 and E2 in Fig. 4a and Table 2). It should be noted that no DBPs were measured in the oxidation of iopamidol by chlorine dioxide (sequence A3 in Figs. 4a, b and Table 2). Our preliminary experimental results also discovered that no reactions took place between these two reagents. But the photolytic products by UV pretreatment did generate certain DBPs with chlorine dioxide. For example, when the applied UV fluence was 1788 mJ cm
2, considerable amount of I-DBPs (including CHCl2I (6.4 mg L
1), CHClI2 (24.0 mg L
1), CHI3 (148.6 mg L
1) and CI3COOH (13.3 mg L
1)) and a trace of DCAN (3.2 mg L
1) were detected (sequence E3 in Figs. 4a, b and Table 2). The increase of UV fluence followed by sequential Cl2, NH2Cl and ClO2 oxidations increased not only increased I-DBP yields but also the species of I-DBPs (Fig. 4b). Direct oxidation by chlorine or monochloramine produced similar amounts of I-DBPs (sequences A1 and A2 in Fig. 4b and Table 2). Nevertheless, an overwhelming majority of CHCl2I with small amount of CHClI2 were formed during chlorination (sequences A1 in Fig. 4b and Table 2) while CHI3 was dominant during chloramination (sequences A2 in Fig. 4b and Table 2). The yields of I-THMs were higher than those in the previous study by Duirk et al. (2011), which resulted from the different experimental conditions. Besides, no HAAs were detected during both Cl2 and NH2Cl oxidation processes (Fig. 4).

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With increased UV fluence, the distinct higher I-DBPs formation was observed during monochloramination than during chlorination (Fig. 4b and Table 2), which was in agreement with literature reported by other researchers (Bichsel and von Gunten, 2000; Richardson and Postigo, 2012; Smith et al., 2010). Photodecomposition of iopamidol by UV resulted in the release of I
in solution (Doll and Frimmel, 2004; Jeong et al., 2010), which could be oxidized to HOI by chlorine, monochloramine and chlorine dioxide. The organic by-products originated from iopamidol UV degradation then reacted with the secondary oxidant, HOI, to form I-DBPs. But the next reactions from I
to form IO
3 with chlorine are much faster than reactions to form other I-DBPs, while the corresponding re
actions with chloramine (to form IO
2 and IO3 ) are much slower, such that I-DBPs increased in formation in the later process (Bichsel and von Gunten, 1999, 2000; Richardson and Postigo, 2012). On the other hand, UV irradiation followed by ClO2 oxidation of iopamidol formed significant amount of CHI3 (sequences B3, C3, D3 and E3 in Fig. 4b and Table 2), which is consistent with our previous study on the oxidation of iodide-containing waters with ClO2 (Ye et al., 2013). I
in solution from deiodination of iopamidol was rapidly oxidized to I2 in the presence of ClO2 (Fabian and Gordon, 1997). Then disproportionation of I2 leads to the formation of HOI (Ye et al., 2013). The indirect transformation routes greatly suppressed production of HOI, which might account for less HOI than chlorination and chloramination, so that less I-DBPs were formed. As presented in Table 2, the conversion of iodine to IDBPs displayed a uniform tendency as discussed above, which increased in the order ClO2 < Cl2 < NH2Cl. The organoleptic threshold concentration of CHI3, the predominant I-DBP formed in this study, lies between 0.03 and 1 mg L
1, which is the lowest of all the I-THMs (Bichsel and von Gunten, 2000). Accordingly, it should be noted that the CHI3 formation in the oxidation after UV treatment of waters containing certain iopamidol may cause serious toxic and taste or odor problems, especially during chloramination. The risk of production of CHI3-related problems in UV pretreated waters decreases in the order NH2Cl > Cl2 > ClO2. It has been suggested that for Cl2 and NH2Cl, which can react with iopamidol, direct formation is an important pathway to generate I-DBPs (Duirk et al., 2011). Therefore, UV deiodination of iopamidol followed by oxidation might be an unnegligible mechanism in formation of I-DBPs as well. It can also be concluded that the pretreatment of iopamidolcontaining waters by UV irradiation enhanced or even shifted the formation of DBPs to more iodinated species from subsequent disinfection processes. The results from the current study lead to a few practical concerns in the application of UV and Cl2, NH2Cl or ClO2 sequential disinfection.

4.

Conclusions

UV irradiation was proved to be effective for iopamidol degradation with first-order kinetics and the quantum yield was evaluated as 0.03318 mol einstein
1. Iopamidol with lower concentration could be degraded more effectively by UV

irradiation, while the effect of pH on the UV degradation rate was negligible. UV irradiation induced the stepwise deiodination of iopamidol. The destruction pathways of iopamidol by UV irradiation based on the products identified by the analyses of UPLC-ESI-MS was proposed and the iodide released was detected. The TOC concentration maintained relatively stable throughout the UV irradiation of iopamidol, indicating no mineralization to CO2. The increase of UV fluence depressed the formation of classical-DBPs while significantly enhanced the formation of I-DBPs during sequential oxidation by Cl2, NH2Cl or ClO2. It was supposed that the oxidation of iodide, which derived from UV deiodination of iopamidol, motivated the striking results. Therefore, concerns should be drawn for the application of UV and sequential disinfection processes by Cl2, NH2Cl and ClO2 in DWTPs.

Acknowledgments This study was supported in part by the Natural Science Foundation of China (No. 51278352 and 41301536) in China, the Fundamental Research Funds for the Central Universities, the National Major Science and Technology Project of China (No. 2012ZX07404 and 2012ZX07408001) and the National Science Council of Taiwan (NSC-102-2221-E-327-021). The authors would like to thank the anonymous reviewers for their valuable comments and suggestions to improve the quality of the paper.

Appendix A. Supplementary data Supplementary data related to this article can be found at http://dx.doi.org/10.1016/j.watres.2014.03.069

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