key: cord-0699927-dwht8vjx
authors: Tian, Fu-Xiang; Ye, Wen-Kai; Xu, Bin; Hu, Xiao-Jun; Ma, Shi-Xu; Lai, Fan; Gao, Yu-Qiong; Xing, Hai-Bo; Xia, Wei-Hong; Wang, Bo
title: Comparison of UV-induced AOPs (UV/Cl(2), UV/NH(2)Cl, UV/ClO(2) and UV/H(2)O(2)) in the degradation of iopamidol: Kinetics, energy requirements and DBPs-related toxicity in sequential disinfection processes
date: 2020-05-30
journal: Chem Eng J
DOI: 10.1016/j.cej.2020.125570
sha: 624e03e020fef7813bcf91d01de342dfc5988bc7
doc_id: 699927
cord_uid: dwht8vjx

The UV-induced advanced oxidation processes (AOPs, including UV/Cl(2), UV/NH(2)Cl, UV/ClO(2) and UV/H(2)O(2)) degradation kinetics and energy requirements of iopamidol as well as DBPs-related toxicity in sequential disinfection were compared in this study. The photodegradation of iopamidol in these processes can be well described by pseudo-first-order model and the removal efficiency ranked in descending order of UV/Cl(2)>UV/H(2)O(2)>UV/NH(2)Cl>UV/ClO(2)>UV. The synergistic effects could be attributed to diverse radical species generated in each system. Influencing factors of oxidant dosage, UV intensity, solution pH and water matrixes (Cl(-), NH(4)(+) and nature organic matter) were evaluated in detail. Higher oxidant dosages and greater UV intensities led to bigger pseudo-first-order rate constants (K(obs)) in these processes, but the pH behaviors exhibited quite differently. The presence of Cl(−), NH(4)(+) and nature organic matter posed different effects on the degradation rate. The parameter of electrical energy per order (EE/O) was adopted to evaluate the energy requirements of the tested systems and it followed the trend of UV/ClO(2)>UV>UV/NH(2)Cl>UV/H(2)O(2)>UV/Cl(2). Pretreatment of iopamidol by UV/Cl(2) and UV/NH(2)Cl clearly enhanced the production of classical disinfection by-products (DBPs) and iodo-trihalomethanes (I-THMs) during subsequent oxidation while UV/ClO(2) and UV/H(2)O(2) exhibited almost elimination effect. From the perspective of weighted water toxicity, the risk ranking was UV/NH(2)Cl>UV/Cl(2)>UV>UV/H(2)O(2)> UV/ClO(2). Among the discussed UV-driven AOPs, UV/Cl(2) was proved to be the most cost-effective one for iopamidol removal while UV/ClO(2) displayed overwhelming advantages in regulating the water toxicity associated with DBPs, especially I-THMs. The present results could provide some insights into the application of UV-activated AOPs technologies in tradeoffs between cost-effectiveness assessment and DBPs- related toxicity control of the disinfected waters containing iopamidol.

At present, due to the worldwide outbreak of atypical pneumonia, caused by COVID-19, which is limitedly known to be sensitive to UV light and effectively killed by chlorine-based disinfectants，disinfection has never been valued so seriously by everybody [1] [2] . Some investigators have proposed that COVID-19 might have potential risks of water mediated transmission, which presents much stricter requirements and objectives for water disinfection [2] . However, more than 600 kinds of disinfection by-products (DBPs) have been accidentally found when chemical disinfectants are used to kill aquatic pathogens. The vast majority of DBPs have not been chemically or biologically characterized and only less than 100 kinds were resolved in quantitative occurrence or toxicity research centers [3] . DBPs are inevitably formed when commonly used disinfectants (Cl 2 , O 3 , ClO 2 or NH 2 Cl, etc) react with naturally occurring organic matters in source waters during disinfection [3] [4] . 11 types of DBPs, such as trihalomethanes (THMs) and haloacetic acids (HAAs), have been controlled in USA while 74 emerging DBPs are not regulated because of their medium occurrence levels or toxicological properties [3] [4] . Besides, primary treatment with disinfectants is often followed by post disinfection to maintain disinfectants residues in water supply systems [5] [6] . The occurrence of some emerging DBPs (for example, nitrogen-containing DBPs, iodo-DBPs (I-DBPs)) with highly enhanced toxicity might be unavoidable in this process [3, 6] .

The comparative toxicity order of different DBPs categories is iodination˃ bromination˃chlorination [7] . This indicates that the more toxic I-DBPs are more worthy of studying on the health risk assessment in drinking water disinfection [3, 8] .

Among I-DBPs, which have attracted more and more attention nowadays, iodoacetic acid (IA) is an emerging drinking water pollutant with high toxicity. Studies have -4- shown that IA is the most genotoxic DBPs in mammals [9] [10] . Iodo-THMs (I-THMs) are the most detected I-DBPs species in drinking water. In addition to their far higher toxicity than conventional THMs, I-THMs could also cause other problems to potable water [11] . For example, iodoform has been considered as an important contributor to the incidents of bad tastes and odors, owing to its lowest organoleptic threshold concentration (0.03-1μg L -1 ) in all THMs [12] . In fact, it was estimated that up to 25% of bad taste and odor events in French drinking water were caused by iodoform [13] [14] . Therefore, the regulation of I-DBPs is of great significance in lowering drinking water toxicity during disinfection. To achieve this goal, control measures on the transformation of iodine sources are extremely practical and operable.

Iopamidol is proved to be the most important organic iodine source of I-DBPs in drinking water [15] [16] . The UV transformation characteristics of iopamidol to I-DBPs in subsequent oxidation processes have been reported in our previous work [15] . The degradation features and pathways of iopamidol in three UV-based advanced oxidation processes (AOPs, UV/Cl 2 , UV/S 2 O 8 2- and UV/H 2 O 2 ) were also studied [17] [18] . Some investigators demonstrated that processes of O 3 /H 2 O 2 and UV/TiO 2 could also remove iodinated X-ray contrast media (ICM) including iopamidol [19] [20] .

Wang et al. studied the UV/Cl 2 degradation of iohexol (another important ICM) and

I-THMs formation in the following chlorination [21] . Despite of so many reports on ICM degradation, most works concentrated on influence factors and destruction mechanisms in UV/Cl 2 and UV/H 2 O 2 , while UV/NH 2 Cl and UV/ClO 2 have never been considered yet. Although the degradation of iopamidol by UV/Cl 2 and UV/H 2 O 2 have been reported [17] [18] , the required electrical energy are still unknown and no effective measures haven been proposed to restrain its transformation to noxious I-DBPs. Thus, this study aimed to make overall comparisons of the UV-induced AOPs -5- with respect to degradation characteristics, energy consumption and toxicity evaluation, especially the never touched processes of UV/NH 2 Cl and UV/ClO 2 .

Recently the UV-initiated AOPs, which combine UV irradiation with common oxidants (such as Cl 2 , NH 2 , HSO 5 -) or synthesized photocatalysts have attracted increasing interests in water treatment and environmental remediation fields [17] [18] [22] [23] [24] [25] . Various highly reactive radicals can be formed in these integrated systems, which can apparently promote the degradation rate of organic contaminants compared with UV alone [17] [18] [24] [25] [26] . Research on iohexol indicated that UV/Cl 2 had more advantages than UV in controlling I-THMs [21] . Other studies on UV/Cl 2 also showed that higher Cl 2 concentrations could not only degrade ICM more effectively but also reduce the formation of I-THMs [27] [28] [29] . The UV photolysis of NH 2 Cl can form Cl• and NH 2 • while the former could directly or indirectly degrade many pollutants by forming •OH [30] [31] [32] . Although UV/NH 2 Cl might inhibit the degradation of organic compounds and promote the DBPs formation in actual waters, the removal effect was fairly satisfactory [23, 33] . H 2 O 2 can also be UV photolyzed directly to form •OH and the highly reactive •OH is considered to play the most important role as it could attack organic compounds by inducing a series of oxidation reactions [34] [35] [36] [37] [38] . However, the evolution of iopamidol in these UV-based AOPs and their effects on the formation of I-DBPs in subsequent oxidation are still not clear. Ultra-pure water generated from a Milli-Q water purification system (Millipore, USA)

were used to prepare all the solutions here.

NH 4 Cl and NaOCl (molar ratio of Cl 2 /N=0.8:1) at pH 8.5 were newly mixed to prepare NH 2 Cl solution [39] . Stock solution of ClO 2 was generated freshly to ensure the purity using the modified method detailed previously [40] . The water matrix of real water was used as NOM sources in this study. The real water samples were acquired from the river running through our university and purified by 0.45 μm membranes (Millipore Corp., Billerica) before used. The main water characteristics of real waters were given in Table S10 .

The experiments on the degradation of iopamidol by different UV-based processes (UV alone, UV/Cl 2 , UV/NH 2 Cl, UV/ClO 2 and UV/H 2 O 2 ) were carried out in a photoreactor equipped with four Hg UV lamps of low pressure (TUV11WT54P-SE, Philips, Netherlands) described elsewhere in our former reports [15, 41] . In DBPs (including classical DBPs and I-THMs) formation experiments, the iopamidol solution spiked with certain dose oxidant (Cl 2 , NH 2 Cl, ClO 2 and H 2 O 2 ) was exposed to the same UV fluence to realize a complete transformation of iopamidol.

Then certain volume of Na 2 S 2 O 3 was added, which was calculated based on the residual concentration of oxidant measured after the photochemical reaction.

Afterward, triplicate samples in 40-mL glass under headspace-free conditions were then placed into an incubator, which was kept in dark for 25 o C inside. After being incubated for 7 days, the samples were quenched with NH 4 Cl (20% in excess) and then were extracted by MtBE for classical DBPs and I-THMs analysis, respectively.

Analytical informations of the chemicals (including iopamidol, classical DBPs and I-THMs) detected in this study were presented in Table S1 .

The analysis of iopamidol was achieved by HPLC (Shimadzu LC-20A) with a XTerra® MS C 18 column (4.6×250mm i.d., 5μm film thickness, Waters, USA) and an UV detector at wavelength of 242nm [15] . Volume ratio of 1:9 between acetonitrile and water and flow rate of 0.80mL min -1 was used. The analytic limitation was 10.0µg L -1 and 10µL of the injection volume was intercalated for detection.

The analytical methods of classical DBPs and I-THMs were modified based on USEPA Method 551.1 [15, 42] . The treated samples were extracted by MtBE and then analyzed by a gas chromatograph (GC-2010, Shimadzu, Japan) fixed with an electron capture detector and a HP-5 capillary column (30m×0.25mm i.d., 0.25µm film thickness, J&W, USA). The oxidized samples were adjusted to pH 5-6 before detecting total organic carbon (TOC) as well as total nitrogen (TN) and then analyzed by TOC-L (Shimadzu, Japan) with ASI-L auto-sampler. The detection limit of TOC was 0.1 mg L -1 . UV 254 was detected by an UV-Vis spectrophotometer (SQ-4802 UNICO, Shanghai) via a 1-cm quartz cell. H 2 O 2 was quantified by the titanium oxalate method spectrophotometrically [43] . The concentrations of Cl 2 and NH 2 Cl were both detected by the N, N-diethyl-p-phenylenediamine (DPD) method [44] . University, Beijing, China) placed into a quartz sleeve inside the reactor was used to measure the exposed light intensity, which were altered by turning on 1, 2, 3 and 4 lamps (measured to be 2.43, 4.94, 7.34 and 9.76 mW cm -2 , respectively).

The degradation of pollutants by UV/Cl 2 [17-18, 26-27, 45-46] , UV/NH 2 Cl [30, [47] [48] [49] , UV/ClO 2 [50] [51] [52] and UV/H 2 O 2 [53] [54] [55] [56] processes are developed and the kinetics could be depicted by pseudo-first-order model with regard to the compound concentration, as explicated in equation (1) .

The following equation (2) can be acquired by integrating equation (1):

Where obs k represents rate constant of the pseudo-first-order reaction, t is the corresponding reaction time, C 0 and C t stand for the initial and instant concentrations of iopamidol. The total decomposition of iopamidol can be attributed to the degradation contribution of reactive radicals ( radicals k ), UV irradiation ( UV k ) and the oxidant itself ( oxidant k ), respectively, viz., [17, 30, 47, 51, 56] . Our previous studies have reported the oxidation rate constants of iopamidol by Cl 2 and NH 2 Cl, which revealed the negligible degradation of the target compound after 300 s reaction (the same data shown in Fig. S1 ) [57] . It was also found that both H 2 O 2 and ClO 2 can not oxidize iopamidol evidently (data also shown in Fig. S1 ) within 300 s [17] [18] . Hence the values of oxidant k in the respective system can be neglected. Therefore, the obs k can be simplified as: (4)). The UV degradation of iopamidol has also been proposed to follow pseudo-first order kinetics [15] . To compare these UV-induced

AOPs, the degradation of iopamidol was thus evaluated firstly, as depicted in Fig.1 .

As presented in Fig UV/Cl 2 exhibited the highest oxidation performance towards iopamidol, due to the comprehensive action of UV photons, the primary radicals (•OH and Cl•) as well as the secondary radicals (Cl 2 • and ClO•), especially the aftermost [17] [18] . It is reported that the UV photolysis rate of chlorines is much higher than that of H 2 O 2 [18, 29, 46, 58] . The highly selective RCS were supposed to be the major contributors while UV photolysis constituted only 13.055% during the UV/Cl 2 degradation of iopamidol. The quenching experiments applying BA and TBA were performed (Text S1 and Figs.S3 (a)-(b)), which indicated that the major contributing radicals were Cl 2 • and ClO•. The results were consistent with the reported studies [17] [18] .

The major contributing radical in UV/H 2 O 2 was believed to be •OH [18, 35, 56] and this can also be demonstrated by the quenching experiments of TBA (Text S1 and Fig.S3 (f)). It was calculated that •OH contributed 80.82% of the total iopamidol degradation ( Fig.1) , the reaction rate constant of which reported to be (3.42±0.28)× 10 9 M -1 s -1 [59] . NH 2 Cl, which slowly releases HClO, would decompose by UV to produce Cl• and NH 2 • (reaction (1) in Table S3 ) [47] , and Cl• will then continue to produce •OH (reactions (2)-(4) in Table S3 ) [30] . Although the direct oxidant was NH 2 Cl, it was the indirect oxidants of free chlorine (HOCl and OCl -) that motivated the following oxidation of iopamidol. Cl• and •OH were suggested to be the major contributing radicals in UV/NH 2 Cl responsible for the degradation of iopamidol [48] [49] , with UV photolysis accounted for only 23 Table S4 ) were too limited to present a high obs k value. Scavenger study using TBA stated that •O was the major contributing radical while other radicals made minor contributions (Text S1 and Fig.S3 (e)). There's hardly any study inquiring into the radical speciations of UV/ClO 2 and more work should be done to distinct the respective contributions of radicals involved.

Simply from the degradation efficiency point of view, UV/Cl 2 technology was the first choice for the removal of iopamidol while UV/ClO 2 was the most reluctant one.

The degradation rate in UV/Cl 2 was up to 4.45 times faster than that in UV/ClO 2 . As for the application convenience, UV/Cl 2 also displayed great advantages on prices and accessibility of chemicals, specially considering the in situ preparation of ClO 2 in water works. Of course, more influence factors other than degradation efficiency should be taken into account in applying these UV-induced AOPs. S4-S11 applying the pseudo-first-order kinetics discussed in Section 3.1. The calculated obs k , t 1/2 (half-life) and the determination coefficient (R 2 ) were shown in Figs. S4-S11 and Tables S6-S9 for comparison. Each curve was comprised of at least six experimentally data and the respective R 2 values were relatively high (R 2 >0.95) to

give enough analytical accuracy. The data might provide useful referential -14- information for applications of the UV-mediated AOPs technologies relating to operational parameters as oxidant dosage, UV intensity, pH and water matrixes.

The UV/Cl 2 degradation of iopamidol with different influencing factors was plotted in Figs. 2 and S4-S5. The reactive radicals and related reactions included in UV/Cl 2 were summarized in Table S2 . obs k , t 1/2 and iopamidol degradation after 300 s reaction under different conditions were provided in Table S6 . were responsible for the degradation of iopamidol at pH 7 [18] . The obs k rank of pH behavior was pH 7 > pH 8 > pH 6 > pH 9 > pH 5. Solution pH can greatly affect the dissociation of HOCl (reaction (6) in Table S2 ). The formations of Cl• and •OH were reduced as pH increased because of the weakened photolysis of chlorines and the enhanced radical scavenging [27] . The decreased obs k as pH fluctuated around 7 can be ascribed to the combined consequence of the declined radical productions and the simultaneous consumption of •OH by OH -(reaction (3) in Table S2 ) [18] . 

The UV/NH 2 Cl degradation of iopamidol with different influencing factors was shown in Figs. 3 and S6-S7. The involved radicals and related reactions included in UV/NH 2 Cl were sorted in Table S3 . obs k , t 1/2 and iopamidol degradation after 300 s reaction under different conditions were provided in Table S7 .

As shown in Figs. 3 (a) and S6 (a), the synergistic degradation of iopamidol may result from the formation of •OH and •Cl in solution [23] , and the productions of these species were promoted as the concentrations of NH 2 Cl added. However, the relevant rate growth slowed down obviously. This could be attributed to the transformation of primary radicals to nitrogen-containing radicals (NH 2 • and NHCl•, etc), which was considered to have lower oxidative capacity and could be ignored in the degradation of iopamidol (reactions (2), (5) (10) and (12) in Table S3 ) [23, 30, 63] .

Moreover, chloramines could also eliminate •OH and the reaction was more easily to occur than that between •OH and HClO (reactions (5) and (12) in Table S3 ) [30, 64] .

The results were consistent with previous studies [23, 31] .

The data in Fig. 3 (b) showed that the degradation rate of iopamidol was distinctly improved as UV intensity added. As depicted in Fig. S6 ( intensified. The dual action upon iopamidol resulted in the higher obs k . The radicals contributed as high as 76.35% to the total degradation in UV/NH 2 Cl (Fig.1) . The increase of UV intensity provided much more reactive radicals and UV photons to iopamidol degradation. The results agreed well with former reports [30, 65] .

The pH behavior of UV/NH 2 Cl degradation upon iopamidol was provided in Figs. 3 (c) and S6 (c). It was observed that solution pH also displayed a significant effect on the degradation rate. The highest obs k in the pH range of 5-9 appeared at pH 7, which was consistent with that of UV/Cl 2 system. This might be ascribed to the homologous radicals of •OH and •Cl in both processes [32, [66] [67] [68] . obs k followed the order of pH7>pH6>pH5>pH8>pH9 and the tendency was slightly different from that in UV/Cl 2 . The phenomena may be explained by the self-degradation of NH 2 Cl and the equilibrium transfer of various chloramine forms (reactions (13)- (16) in Table S3) during the variation of pH [69] [70] . Besides, OH could consume •Cl at higher pHs -18-more quickly (reaction (2) in Table S3 ) [16] . Our previous reports have confirmed that the chloramination of iopamidol was also based on the action of free chlorines (HOCl and ClO -, especially the more reactivie ClO towards iopamidol) from the NH 2 Cl dissociation [57] . Under alkaline conditions, the self-degradation of NH 2 Cl (reaction (16) in Table S3 ) shifted to the right, resulting in great inhibition to reaction (15) in Table S3 . At the acidic pHs, the main chloramine species were NHCl 2 and reaction (13) in Table S3 shifted leftward, thus the obs k were higher than those in alkaline solutions. The consequences were caused by both transformation of chlorine species and equilibrium of HOCl and ClO -. This may be interpreted by the most production of ClO at neutral pH, while at pH 5-6 and pH 8-9, the conversion to ClO was greatly restrained, so the obs k decreased obviously.

The effect of water matrixes including Cl -, NH 4 + and NOM on the UV/NH 2 Cl degradation of iopamidol was plotted in Fig. S7 . As depicted in Fig. S7 

The degradation of iopamidol by UV/ClO 2 with different influencing factors was shown in Figs. 4 and S8-S9. The reactive radicals and related reactions included in UV/ClO 2 were listed in Table S4 . obs k , t 1/2 and iopamidol degradation after 300 s reaction under different conditions were provided in Table S8 .

As shown in Figs. 4 (a) and S8 (a), obs k increased with the increasing ClO 2 concentrations from 0 to 1000 μM due to the added yields of radicals. Although

UV-based AOPs, the roles of radicals were also nonnegligible in this system. The formation of •OH has been proposed in UV/ClO 2 assisted decolorization of methylene blue as well as color and chemical oxygen demand removal [50, 52] . As listed in (7)) [75] . But according to the data here, the radical yields were too low to present an outstanding promotion on the degradation rate. The pH impact on obs k was also evaluated in the range of 5-9 and displayed in Figs. 4 (c) and S8 (c). The degradation rate gradually increased with solution pH but the obs k added limitedly with pH ascending from 7 to 9. The higher obs k were noticed under neutral and basic media at pH 7-9 but decreased in acidic media at pH 7-5. Our previous study has confirmed that pH can not affect the UV photolysis rate of iopamidol and there is no dissociated form of iopamidol at all investigated pHs [15] .

Such pH behavior here might be the integrated degradation results of multi-radicals as 

The findings stated that the removal of iopamidol can be achieved by the combination of UV and ClO 2 and the degradation rate was favorable to higher 

The degradation of iopamidol by UV/H 2 O 2 with different influencing factors was shown in Figs. 5 and S10-S11. The reactive radicals and related reactions included in UV/H 2 O 2 were summarized in Table S5 . obs k , t 1/2 and iopamidol degradation after 300 s reaction under different conditions were provided in Table S9 .

As depicted in Figs. 5 (a) and S10 (a), the data suggested that the elimination rate of iopamidol ( obs k ) gradually increased with added H 2 O 2 dosage but the relationship was not linear. The results can be ascribed to the increased radicals ( (Fig.1) . The joint action of both aspects led to the promoted obs k at higher UV intensities. The results agreed broadly with previous studies by other scholars [54, 56] .

The pH effect on the UV/H 2 O 2 degradation of iopamidol was plotted in Figs. 5 (c) and S10 (c). The obs k distinctly decreased with pH increasing from 5 to 9. The increasing OH in alkaline conditions caused the consumption of •OH and the simultaneous formation of O -• (reactions (9) and (10) in Table S5 ) [78] . It is supposed that •OH acts as electrophile while O -• behaves as nucleophile in their respective reaction with organic molecules [79] . So as pH increased, the less reactive and more selective radical of O -•, rather than •OH, became the major radical towards iopamidol [59, 79] , which then displayed a tardy degradation rate. Besides, the I originated from the UV deiodination also slightly inhibited the removal of iopamidol [17] . These dual effects resulted in a distinct decrease of obs k under basic pHs. The pH behavior here was in correspondence with other relative reports [17] [18] .

The effect of water matrixes of Cl -, NH 4 + and NOM on the UV/H 2 O 2 degradation of iopamidol was given in Fig. S11 . It was seen from Fig. S11 can be applied flexibly under full cost-effective assessment.

As discussed above, the removal effectiveness and major factors of different UV-driven AOPs on iopamidol were investigated comprehensively. As all these photodegradation processes are electric-energy-intensive, it is necessary to introduce a scale-up parameter correlated with the electric energy, which can represent a primary fraction of the operating costs with regard to these techniques [25, 80] . Therefore, the

proposed by the international union of pure and applied chemistry (IUPAC) was then employed to provide a cost-effective evaluation for the concerned systems [81] . O E E / (kWh m -3 order -1 ) is defined as the electrical energy (kWh) required to decompose the contaminants by one order magnitude in 1 m 3 of polluted water [25, [81] [82] 

Where P represents power of the electronic energy input of the UV device (kW, as listed in Text S2); V is the volume of the solution (L); t is the photodegradation time Fig. S2 and Tables S6-S9.

As shown in Fig. S2 

increases, the energy efficiency of a system decreases [84] . UV/Cl 2 was observed to be the most cost-effective one for iopamidol degradation among the tested AOPs. Contrarily, UV/ClO 2 was the highest energy consumption system owing to the biggest equivalent electrical energy of ClO 2 and the lowest degradation rate. Compared with direct UV photolysis,

decreased obviously in the discussed processes due to the contribution of various radical species in obs k .

which proved the significant role of UV intensity in all of these UV-mediated AOPs.

As presented in Tables S6-S9, the necessary UV energy (

reached its maximum in each process. The results of current study can provide a few usefully cost-effective concerns in the application of these UV-driven AOP technologies.

It can be concluded from the above results that the four UV-induced AOPs were all capable of degrading iopamidol effectively compared with UV alone. Our previous research has confirmed the formation of I released from iopamidol and its intermediates during UV irradiation, which then resulted in an enhanced conversion of classical DBPs to highly toxic I-DBPs during subsequent oxidation [15] . Other scholars have also verified deiodination as the vital pathways and the evolution of inorganic iodine (I and HOI) in the degradation of iopamidol by UV/H 2 O 2 and UV/Cl 2 [17] [18] . The I present in these UV based systems can be oxidized to HOI, which further reacted with degradation intermediates of iopamidol and probably led to the formation of undesirable I-DBPs [13, 15, 17] .

The organic oxidation products of iopamidol by UV, UV/H 2 O 2 and UV/Cl 2 have been identified and the destruction mechanisms mainly include deiodination, hydroxylation, chlorination and H-abstraction [15, [17] [18] . The degradation intermediates of iopamidol by UV/Cl 2 , UV/NH 2 Cl, UV/ClO 2 and UV/H 2 O 2 were also analyzed here and the results were shown in Table S11 and Fig. S12 . It was seen that the deiodinated and hydroxylated products were all detected in the four processes. The chlorinated products were both observed in UV/Cl 2 and UV/NH 2 Cl, while chlorine substitutions and hydroxy substitutions of iodine atoms might simultaneous take place, which resulted in much more intermediates than other two systems. The detected products of UV/ClO 2 and UV/H 2 O 2 were similar except for slight monochlorinated products (products 8 and 12 in Table S11 ). It should also be pointed out that amino group substituted products (such as product 2 in Table S11 ) were noticed in UV/NH 2 Cl, which might pose serious influence on the formation of nitrogencontaining DBPs. These oxidized intermediates can provide sufficient organic carbon source and may then facilitate the subsequential attack by oxidants to form DBPs.

The main parameters (TOC, TN and UV 254 ) of iopamidol solutions after treated by different processes were presented in Table S10 . It can be seen that removals of were also calculated in Table S12 .

As seen from the corresponding impact on the toxicity of disinfected waters should not be ignored due to their higher toxicity than carbon-containing DBPs [85] . The attack of radicals in these systems led to the cleavage of asymmetric side chain at the amide bond of the hydroxylated products of iopamidol, and then formation of nitrogen-containing DBPs in the following oxidation [17] . The contribution of a certain DBP to the water toxicity after disinfection depends on its concentration as well as the toxic potency [86] . In order to better estimate the effects of various UV-combined AOPs on the toxicity of disinfected waters, the toxicity weighted concentrations of all tested DBPs were calculated and presented in Fig. 6 (c) and Table S12 .

In these five UV-activated systems, I could be released by UV deiodination of iopamidol [15] . However, the following reactions between I and various reactive species might result in the evolution of iodine speciation and thus affected the final formation of I-THMs in subsequent oxidation. The reaction rates of ROS and RCS -29- towards iopamidol were quite different [17] [18] 87] , and RCS were noticed to be more selective and efficient [17] . Consequently, the relevant products in subsequent reactions with I and the deiodinated intermediates of iopamidol differed greatly. The chlorinated intermediates of iopamidol by RCS were more liable to form I-THMs. On the contrary, ROS were more reluctant to induce I-THMs formation than RCS in the subsequent oxidation not only due to its reaction with I as reactions (22)-(23) but also the formation of hydroxylated products [79, 87] . Table S12 . UV/Cl 2 led to the greatest yields of classical DBPs, which may be attributed to its strongest oxidation ability towards iopamidol. This can also be evidenced by the fastest degradation rate (Fig. 1 ) and the most reduction of TOC as well as UV 254 in UV/Cl 2 ( [79, 87] , hence almost no I-THMs were detected in subsequent chlorination and chloramination (F 1 and F 2 in Fig. 6 ). It was probably due to the reactions of Cl 2

and NH 2 Cl with the iodine species as HOI and I 2 . It should also be pointed out that certain amounts of TIM was detected in F 3 , which may owe to the iodine active substances (HOI, I 2 and I 3 -) produced by the subsequent ClO 2 oxidation (reactions (23) , (27)- (28)) [79, 88] . Besides, there were scarcely any I-THMs formed in UV/ClO 2 . The O 3 produced by reaction (4) in Table S4 can easily oxidize I to IO 3 -(reaction (29)), thus the formation of I-THMs were completely restrained [11] . The iodine conversion of iopamidol to I-THMs ( can provide referential value for the application of chlorine-based disinfectants.

It can also be seen from Fig. 6 that among three subsequent processes, ClO 2 produced the least DBPs, thus presented the smallest water toxicity. The data provided some evidences that ClO 2 could be a green disinfectant in controlling DBPs as Cl 2 and NH 2 Cl alternatives [93] . ClO 2 alone could not oxidize iopamidol and produced none DBPs [15] , but the pretreatment of UV-based processes induced a certain amount of DBPs. The iodine conversion to I-THMs followed the order of UV/NH 2 Besides, it should be specially pointed out that UV/ClO 2 can degrade iopamidol effectively with hardly any I-THMs formed during subsequent oxidation.

Up to present, iopamidol is reported to be the most important organic iodine source for I-DBPs [15] [16] 94] and the fate of iodine in iopamidol has critical significance in evaluating the relevant toxicity of waters. It was concluded that during the subsequent chlorination and chloramination, the iodine conversion in UV/NH 2 Cl turned out to be the highest while those in UV/ClO 2 were quite low. As for the I-THMs-related safety in subsequent disinfection, the advantage of ClO 2 over Cl 2 and NH 2 Cl was obviously noticed. The results can provide application value for these UV-induced techniques in treatment with iopamidol-contaminated waters as well as the toxicity control relating to I-THMs.

From the perspective of weighted water toxicity after subsequent oxidation, the risk ranking was UV/NH 2 Cl>UV/Cl 2 >UV/H 2 O 2 >UV/ClO 2 . Based on our results, more concerns must be drawn in the utilization of UV/NH 2 Cl and UV/Cl 2 , UV/NH 2 Cl in particular, as these treatments might greatly enhance the water toxicity in the following disinfection. In fact, ammonia nitrogen pollution in surface waters is generally serious while Cl 2 is the most widespread conventional disinfectant in most plants. Hence the safety issue concerning DBPs on the UV combination with NH 2 Cl and Cl 2 should be addressed. Meanwhile, UV/ClO 2 displayed overwhelming advantage in controlling the water toxicity problems associated with DBPs, especially I-THMs, compared with other UV-driven AOPs.

The degradation of iopamidol by four UV-induced AOPs can be described by The results of current study can provide usefully cost-effective concerns and DBPs-related toxicity evaluation in the application of these UV-driven AOP technologies. 

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This study was supported in part by the Natural Science Foundation of China (No. 

☒ The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.☐The authors declare the following financial interests/personal relationships which may be considered as potential competing interests: