磺胺氯哒嗪在海水中的间接光降解

窦琦玮,段佳奇,唐新宇,姚庆祯,苏荣国

磺胺氯哒嗪在海水中的间接光降解

窦琦玮,段佳奇,唐新宇,姚庆祯,苏荣国*

(中国海洋大学化学化工学院,山东 青岛 266100)

以有色溶解有机物(CDOM)作为主要光敏剂研究磺胺氯哒嗪(SCP)间接光降解行为和机理,分析CDOM组成、盐度和pH值对SCP间接光降解的影响.SCP间接光降解速率随CDOM浓度升高而逐渐加快.CDOM产生的光化学反应活性中间体对SCP间接光降解的贡献率不同,其中3CDOM*起主要作用,对SCP间接光降解的贡献率高达77.94%.所用CDOM由4种荧光组分组成,包含3种外源腐殖质(C1,C2,C3)和1种内源腐殖质(C4), SCP间接光降解去除率和3CDOM*浓度分别与荧光组分的相关性大小顺序均为C3>C2>C4>C1.其中C3和C2与[3CDOM*]具有较高线性相关性(R2>0.97),是3CDOM*的主要贡献者.盐度和pH值对SCP间接光降解的影响作用显著.在盐度为15‰时,SCP的间接光降解速率最大.在低盐度范围(0～15‰)内,离子强度效应对间接光降解的促进作用大于无机阴离子带来的抑制作用,使得间接光降解速率随着盐度的升高而加快.溶液pH=(5.00±0.10)时,SCP的间接光降解速率最大.SCP的间接光降解速率随着溶液pH值的升高而减慢,中性和碱性环境不利于SCP的间接光降解.

磺胺氯哒嗪；间接光降解；CDOM荧光组成；环境因素

抗生素磺胺氯哒嗪(SCP)大量用于畜牧业和水产养殖业中的疾病治疗,未被利用的SCP随着污水进入污水处理厂,经处理后入海.但是常用的污水处理方法难以完全去除水体中的SCP,导致SCP在近岸海域的检出率和浓度较高[1].间接光降解是大多数磺胺类抗生素的主要降解途径[2-5].有色溶解有机物(CDOM)是自然水体中主要的光敏剂,其在吸收太阳光后会形成光化学反应活性中间体,主要包括单线态氧(1O2)、羟基自由基(HO·)以及三线态CDOM(3CDOM*)等[6],是促进抗生素降解的重要物质[7].研究[8]指出CDOM的来源会影响其光降解性,同时环境因素如盐度和pH值也会影响水体中部分磺胺类抗生素的光降解[9-11].

目前对于CDOM组成对磺胺类抗生素间接光降解的研究较少,同时海水盐度、pH值等环境因素对于磺胺类抗生素间接光降解的影响尚未可知.因此,本研究选取SCP作为研究对象,以CDOM在SCP间接光降解的作用作为研究重点,考察CDOM产生的活性中间体、不同种类CDOM浓度以及盐度和pH值对于SCP间接光降解的作用.采用激发-发射矩阵光谱(EEMs)与并行因子分析(PARAFAC)结合技术分析CDOM组成和来源, 考察其对SCP间接光降解的影响.以期为准确、有效地预测近海海水中SCP的光化学降解过程和评价SCP带来的生态风险提供科学依据.

1 材料与方法

1.1 材料

磺胺氯哒嗪(SCP):纯度>99%,购于Ark Pharm.称取0.1000g的SCP加入超纯水配制成浓度为1g/L的母液,用0.1mol/L的盐酸和氢氧化钠溶液将母液pH值调节至pH=(8.00±0.10).

CDOM:河腐殖酸(SRHA)(2S101H,国际腐殖酸协会),河富里酸(SRFA)(2S101N,国际腐殖酸协会),河天然有机物(SRNOM)(2S101F,国际腐殖酸协会),腐殖酸(JKHA)(百灵威有限公司).分别称取0.0190g的SRHA、0.0191g的SRFA、0.0195g的SRN-OM和0.0250g的JKHA加入超纯水配成100mgC/L CDOM母液,将所有溶液用0.1mol/L的盐酸和氢氧化钠溶液调节至pH=(8.00±0.10).配制好的溶液均转移到100mL棕色瓶中,避光置于冰箱中4℃冷藏保存.

1.2 光降解实验

1.2.1 不同种类和浓度CDOM影响实验 取13支50mL石英管,按照顺序依次向石英管中加入800.0μL、2.00mL和4.00mL的0.1gC/L的不同种类的CDOM母液(对照管不加CDOM),每一支石英管再分别加入400.0μL的SCP母液并加超纯水定容到40mL,使得配制后的溶液中CDOM浓度分别为2,5,10mgC/L(对照管CDOM浓度为0mgC/L),SCP初始浓度为10mg/L,将所有溶液用0.1mol/L的盐酸和氢氧化钠溶液调节至pH=(8.00±0.10).光降解实验在XPA-7光化学反应仪中进行,光化学反应仪以1000W的氙灯作为模拟自然光源,使用320nm透光膜滤光,反应温度25℃,光照24h,每4h取一次样品,用带有孔径为0.22μm的聚醚砜(PES)膜的注射器过滤后进入1.5mL棕色进样瓶冷藏待测.每组实验均设置3组平行样.

1.2.2 活性中间体影响实验 取4支石英管,分别加入一定量的异丙醇、四氢呋喃、苯酚(1mmol/L,现配现用)储备液和等量的超纯水作为对照,向这4支石英管加入40.00mL的10mgC/L的JKHA储备液以及400.0μL的SCP母液.将光解液溶液用0.1mol/L的盐酸和氢氧化钠溶液调节至pH=(8.00± 0.10).然后将石英管放入光化学反应仪中按1.2.1中的条件进行光解,取样.

1.2.3 盐度影响实验 取5支石英管加入40.00mL的10mgC/L的JKHA储备液以及400.0μL的SCP母液,然后分别加入一定量的标准海水和超纯水(作为对照),使最终配制的40.00mL反应液的盐度分别是5‰、15‰、25‰、35‰,CDOM浓度为10mgC/L,SCP初始浓度为10mg/L,将光解液用0.1mol/L的盐酸和氢氧化钠溶液调节至pH=(8.00± 0.10).然后将石英管放入光化学反应仪中按1.2.1中的条件进行光解,取样.

1.2.4 pH值的影响实验 取4支石英管加入40.00mL的10mgC/L的JKHA储备液以及400.0μL的SCP母液,加超纯水使CDOM浓度为10mgC/L、SCP浓度为10mg/L.然后用0.1mol/L的稀盐酸或氢氧化钠溶液将反应液调节pH值至pH=(5.00±0.10)、pH=(7.00±0.10)、pH=(8.00±0.10)、pH=(9.00±0.10)和pH=(11.00±0.10),用IAPSO标准海水将光解液盐度调节为30‰.然后将石英管放入光化学反应仪中按1.2.1中的条件进行光解,取样.

1.3 光谱分析

待测溶液经孔径为0.22μm的聚醚砜(PES)膜过滤后用荧光分光光度计(Fluorolog3-11)进行荧光测定:激发波长范围240～480nm、发射波长范围250～ 580nm,扫描步长5nm.狭缝宽度5nm,积分时间为0.05s,荧光强度用浓度为10mg/L的硫酸奎宁稀硫酸溶液进行校正定标.

待测溶液经孔径为0.22μm的聚醚砜(PES)膜过滤后用紫外可见分光光度计(UV-2500)进行测定,波长范围为200～800nm,参比液为超纯水.

1.4 液相色谱条件

色谱柱为Agilent PLRP-S(5μm,150×4.60mm) C18液相色谱柱,进样量20μL.流动相为甲酸(0.1%):乙腈=60:40,检测波长为270nm.

2 结果与讨论

2.1 SCP在CDOM溶液中的间接光降解

SCP的最大吸收峰在320nm以下(图1),使用320nm滤光膜过滤掉320nm以下的紫外光,可以有效屏蔽SCP直接光降解.

通过HPLC测定光降解过程中的SCP浓度,以一级反应动力学函数获得间接光降解速率常数obs,2>0.97.

在pH=(8.00±0.10)的条件下,10mg/L的SCP在不同浓度不同来源CDOM溶液中的光降解情况如图2所示.

在添加CDOM后,SCP的间接光降解速率都有不同程度的提高,且随着CDOM浓度的升高,SCP的间接光降解速率均明显加快.其中,JKHA浓度变化的影响最为显著,对于SCP间接光降解的促进作用最强,而SRNOM、SRHA和SRFA对于SCP间接光降解的促进作用基本相等.从4种CDOM的光吸收曲线(图1)可以看出,JKHA的吸光度远高其他3种,对光具有较强吸收作用是JKHA显著促进SCP光降解的主要原因之一.

图2 SCP在CDOM溶液中的间接光降解

CDOM在光化学反应过程中会产生活性中间体,这些活性中间体与有机污染物反应实现有机污染物的间接光降解.常见的活性中间体有3CDOM*、HO·、1O2等,其中3CDOM*除了可以直接降解抗生素以外,还可以通过电子、能量和氢原子转移与水及溶解氧生成HO·、1O2等促进有机污染物光降解[8,12]. Tang等[13]测定了本研究所用4种CDOM溶液中活性中间体的稳态浓度(表1),表明活性中间体的稳态浓度随CDOM浓度的增加而增加.

表1 不同浓度的4种CDOM产生的活性中间体的稳态浓度

2.2 活性中间体影响实验

异丙醇(IPA)通常作为HO·的猝灭剂[14];四氢呋喃(THF)为1O2的猝灭剂[15];苯酚既可以猝灭HO·又可以抑制由3CDOM*引导的光转化过程.本研究采用异丙醇、四氢呋喃、苯酚分别作为HO·、1O2以及3CDOM*的猝灭剂研究SCP的间接光降解机理(图 3).添加苯酚后,SCP的间接光降解速率显著减慢,3CDOM*被猝灭后SCP间接光降解过程受到了明显抑制,说明3CDOM*在CDOM参与的SCP间接光降解中起主要作用[16];在添加四氢呋喃和异丙醇后,SCP的间接光降解速率变化较小,说明1O2和HO·对SCP间接光降解的作用并不显著.

图3 SCP在猝灭剂溶液中的间接光降解

将添加了自由基猝灭剂的obs降低率作为各活性自由基的贡献率(表2),可以看出3CDOM*对SCP间接光降解的贡献率最大,达到77.94%,HO·的贡献率次之,为11.62%,1O2在SCP间接光降解中的作用最小.

表2 各活性中间体对SCP间接光降解的贡献率

2.3 CDOM荧光组成对SCP间接光降解的影响

2.3.1 CDOM荧光组成 CDOM是天然水体中最主要的光敏剂,不同来源的CDOM其组成和化学性质不同,对SCP间接光降解的作用也不同.利用EEMs-PARAFAC解析获得4种CDOM样品的荧光组分C1,C2,C3和C4(图4,表3).

C1的最大激发波长和最大发射波长分别为355和460nm,这与Singh等[17]报道的最大激发/发射波长为370/460nm的陆地源类腐殖质荧光组分相似,也与Coble等[18]报道的激发和发射波长在 350nm/ 420～480nm 的组分峰以及Yamashita等[19]在利物浦湾发现的来自陆地水生环境的微生物衍生的类腐殖质成分(xmax/mmax=(265)370/464nm)相似.因此,C1被认为是一种外源类腐殖质荧光组分.

表3 CDOM 4种组分的荧光强度

C2在300和405nm处出现激发峰值,其中405nm为最大激发波长,最大发射波长为495nm.这与之前报道的最大激发波长和最大发射波长为350～455/492～520nm的来源于土壤的富里酸荧光组分类似[20],也与Yamashita等[21]报道的外源类腐殖质荧光组分(xmax/mmax=390(275)479nm)类似.

C3的两个峰值激发波长分别在335和460nm,其中最大激发波长为460nm,最大发射波长为525nm.这与Singh等[17]报道的组分4(xmax/mmax= 240(410)/520nm)和Lochmuller等[20]报道的“Contech FA”(Exmax/Emmax=390/509nm)相似.该成分被认为是源自农业集水区以及微生物降解等过程中产生的.

C4的最大激发波长和最大发射波长分别为315和425nm,这与之前报道[22]的组分4(xmax/mmax= 325(250)/416nm)和代表海洋类腐殖质的峰值M[23](xmax/mmax=312/380～420nm)相似.该组分被认为是一种内源类腐殖质组分,通常在微生物降解过程中产生.

表4 CDOM荧光组分

在相同的DOC浓度下,CDOM总荧光强度的大小排序为JKHA>SRFA>SRNOM>SRHA(表3),其中JKHA荧光强度明显高于其他3种CDOM,且C2和C3组分占比较高,JKHA中C2和C3荧光强度分别约占总荧光强度的39%和35%.

2.3.2 CDOM荧光组成对SCP间接光降解的影响 C2和C3组分与1O2、HO·和3CDOM*的稳态浓度之间存在较强的相关性(2>0.659)(表5).研究表明外源CDOM的分子量和芳香性均高于内源CDOM[24], C2和C3具有较长的激发和发射波长,分子量大,芳香程度高[25].芳香度高、分子量大的CDOM具有较高的活性中间体产率[26].

4个荧光组分与SCP去除率的相关性及4个荧光组分与3CDOM*浓度的相关性大小顺序均为C3> C2>C4>C1.C2和C3与3CDOM*浓度的相关性较高(2>0.97),表明C2与C3组分是3CDOM*的主要来源.而JKHA中C2与C3组分含量较高,对SCP间接光降解表现出较强的促进作用,这也是JKHA能显著促进SCP间接光降解的原因之一.

表5 SCP间接光降解速率、CDOM组分荧光强度与活性中间体稳态浓度之间的相关性

* 在0.05级别(双尾),相关性显著;** 在0.01级别(双尾),相关性显著.

2.3 盐度对SCP间接光降解的影响

有研究[9-11]表明盐度会抑制水体中抗生素的光降解作用.光解液中JKHA浓度为10mgC/L,光解结果如图5.在实验设置的盐度范围内(0～35‰),随着盐度的升高,SCP的间接光降解速率先加快后减慢,在盐度为15‰时SCP的间接光降解最快.低盐度范围内(0～15‰),盐度的升高对于SCP间接光降解的促进作用逐渐增强,而在高盐度范围内(15‰～35‰),随盐度升高SCP的间接光降解速率越来越慢.

盐度的提高使得离子强度增强,离子强度效应使得海水中3CDOM*稳态浓度显著增加[27],3CDOM*稳态浓度的升高促进了水中SCP的间接光降解.然而,海水中还存在各种无机阴离子,如氯离子和溴离子,无机阴离子可以清除HO·[28-29],从而降低了SCP的间接光降解速率,抑制SCP的间接光降解.所以,在低盐度范围内(0～15‰),离子强度效应对SCP间接光降解的促进作用要大于无机阴离子带来的抑制作用,使得SCP间接光降解速率随着盐度的升高而加快;而在高盐度范围内(15‰～35‰),离子强度效应带来的对SCP间接光降解的促进作用要小于无机阴离子带来的抑制作用,这就导致SCP在高盐度范围内(15‰～35‰)的间接光降解速率随着盐度的升高而减慢.

图5 SCP在不同盐度溶液中的间接光降解

2.4 海水环境下水体pH值对SCP间接光降解的影响

pH值不仅可以影响水体中离子、CDOM以及有机污染物的存在形态,还可以影响活性中间体的生成速率[30],是水体中有机污染物间接光降解的重要影响因素.光解液中JKHA浓度为10mgC/L,光解结果如图6.

SCP的间接光降解速率随着溶液pH值的升高而减慢,在pH=(5.00±0.10)时,SCP的间接光降解速率最大,表明中性和碱性环境抑制了SCP的间接光降解.

改变pH值可以导致可电离化合物的质子化/去质子化,改变化合物的存在形态和吸附性能.SCP的解离常数分别为pKa1=(2.0±0.8)和pKa2=(5.9±0.3),在低pH值条件下,SCP分子发生质子化作用,质子化后的SCP分子可与CDOM产生静电吸附作用,而在中性和碱性条件下,去质子化后的SCP-吸附系数明显低于酸性条件下的SCP+吸附系数,具有更强的静电排斥作用[31],与CDOM的吸附作用变弱,减少了SCP与活性中间体的接触, SCP的间接光降解速率随之减小.

图6 SCP在不同pH溶液中的间接光降解

3 结论

3.1 本文所用4种CDOM对SCP的间接光降解均起促进作用.当CDOM浓度达到10mgC/L时, JKHA对SCP间接光降解的作用远大于其他3种.不同CDOM对SCP间接光降解作用的不同主要是CDOM组成不同和活性中间体产生效率导致的.

3.2 CDOM光解产生的活性中间体如3CDOM*、HO·和1O2等参与SCP间接光降解.其中3CDOM*是SCP间接光降解的主要参与者,其贡献率高达77.94%.

3.3 盐度对SCP间接光降解的影响主要是离子强度效应和卤素离子效应共同作用的结果.在盐度为15‰时,SCP的间接光降解最快.低盐度范围内(0～15‰),盐度的升高对于SCP间接光降解起促进作用.

3.4 水体pH值影响SCP的存在形态,当pH= (5.00± 0.10)时,SCP的间接光降解速率最大,pH值降低,不利于SCP的间接光降解.

[1] Luo Y, Xu L, Rysz M, et al. Occurrence and transport of tetracycline, sulfonamide, quinolone, and macrolide antibiotics in the Haihe River Basin, China [J]. Environmental Science & Technology, 2011,45(5): 1827-1833.

[2] Wei C, Li X, Xie Y, et al. Direct photo transformation of tetracycline and sulfanomide group antibiotics in surface water: kinetics, toxicity and site modeling [J]. Science of the Total Environment, 2019,686: 1-9.

[3] Baena-Nogueras R M, González-Mazo E, Lara-Martín P A. Photolysis of antibiotics under simulated sunlight irradiation: identification of photoproducts by high-resolution mass spectrometry [J]. Environmental Science & Technology, 2017,51(6):3148-3156.

[4] Boreen A L, Arnold W A, McNeill K. Photochemical fate of sulfa drugs in the aquatic environment: sulfa drugs containing five- membered heterocyclic groups [J]. Environmental Science & Technology, 2004,38(14):3933-3940.

[5] Niu X Z, Glady-Croué J, Croué J P. Photodegradation of sulfathiazole under simulated sunlight: Kinetics, photo-induced structural rearrangement, and antimicrobial activities of photoproducts [J]. Water Research, 2017,124:576-583.

[6] Yang W, Abdelmelek S B, Zheng Z, et al. Photochemical transformation of terbutaline (pharmaceutical) in simulated natural waters: Degradation kinetics and mechanisms [J]. Water Research, 2013,47(17):6558-6565.

[7] Larson R A, Zepp R G. Reactivity of the carbonate radical with aniline derivatives [J]. Environmental Toxicology and Chemistry: an International Journal, 1988,7(4):265-274.

[8] McKay G, Couch K D, Mezyk S P, et al. Investigation of the coupled effects of molecular weight and charge-transfer interactions on the optical and photochemical properties of dissolved organic matter [J]. Environmental Science & Technology, 2016,50(15):8093-8102.

[9] Trovó A G, Nogueira R F P, Agüera A, et al. Degradation of sulfamethoxazole in water by solar photo-Fenton. Chemical and toxicological evaluation [J]. Water Research, 2009,43(16):3922-3931.

[10] Yang C C, Huang C L, Cheng T C, et al. Inhibitory effect of salinity on the photocatalytic degradation of three sulfonamide antibiotics [J]. International Biodeterioration & Biodegradation, 2015,102:116-125.

[11] Oliveira C, Lima D L D, Silva C P, et al. Photodegradation of sulfamethoxazole in environmental samples: the role of pH, organic matter and salinity [J]. Science of the Total Environment, 2019,648: 1403-1410.

[12] Niu X Z, Liu C, Gutierrez L, et al. Photobleaching-induced changes in photosensitizing properties of dissolved organic matter [J]. Water Research, 2014,66:140-148.

[13] Tang X, Cui Z, Bai Y, et al. Indirect photodegradation of sulfathiazole and sulfamerazine: Influence of the CDOM components and seawater factors (salinity, pH, nitrate and bicarbonate) [J]. Science of the Total Environment, 2021,750:141762.

[14] Li G, Zhu M, Chen J, et al. Production and contribution of hydroxyl radicals between the DSA anode and water interface [J]. Journal of Environmental Sciences, 2011,23(5):744-748.

[15] Li B, Ahmed F, Bernstein P S. Studies on the singlet oxygen scavenging mechanism of human macular pigment [J]. Archives of Biochemistry and Biophysics, 2010,504(1):56-60.

[16] Guo Z, Wang J, Chen X, et al. Photochemistry of dissolved organic matter extracted from coastal seawater: Excited triplet-states and contents of phenolic moieties [J]. Water Research, 2021,188:116568.

[17] Singh S, D'Sa E J, Swenson E M. Chromophoric dissolved organic matter (CDOM) variability in Barataria Basin using excitation– emission matrix (EEM) fluorescence and parallel factor analysis (PARAFAC) [J]. Science of the Total Environment, 2010,408(16): 3211-3222.

[18] Coble P G, Green S A, Blough N V, et al. Characterization of dissolved organic matter in the Black Sea by fluorescence spectroscopy [J]. Nature, 1990,348(6300):432-435.

[19] Yamashita Y, Panton A, Mahaffey C, et al. Assessing the spatial and temporal variability of dissolved organic matter in Liverpool Bay using excitation–emission matrix fluorescence and parallel factor analysis [J]. Ocean Dynamics, 2011,61(5):569-579.

[20] Lochmueller C H, Saavedra S S. Conformational changes in a soil fulvic acid measured by time-dependent fluorescence depolarization [J]. Analytical Chemistry, 1986,58(9):1978-1981.

[21] Yamashita Y, Jaffé R, Maie N, et al. Assessing the dynamics of dissolved organic matter (DOM) in coastal environments by excitation emission matrix fluorescence and parallel factor analysis (EEM- PARAFAC) [J]. Limnology and Oceanography, 2008,53(5):1900-1908.

[22] Stedmon C A, Markager S, Bro R. Tracing dissolved organic matter in aquatic environments using a new approach to fluorescence spectroscopy [J]. Marine Chemistry, 2003,82(3/4):239-254.

[23] Coble P G, Del Castillo C E, Avril B. Distribution and optical properties of CDOM in the Arabian Sea during the 1995 Southwest Monsoon [J]. Deep Sea Research Part Ⅱ: Topical Studies in Oceanography, 1998,45(10/11):2195-2223.

[24] Richard C, Canonica S. Aquatic phototransformation of organic contaminants induced by coloured dissolved natural organic matter [J]. Environmental Photochemistry Part II, 2005:299-323.

[25] Kalbitz K, Geyer W, Geyer S. Spectroscopic properties of dissolved humic substances—a reflection of land use history in a fen area [J]. Biogeochemistry, 1999,47(2):219-238.

[26] Timko S A, Romera-Castillo C, Jaffé R, et al. Photo-reactivity of natural dissolved organic matter from fresh to marine waters in the Florida Everglades, USA [J]. Environmental Science: Processes & Impacts, 2014,16(4):866-878.

[27] Parker K M, Pignatello J J, Mitch W A. Influence of ionic strength on triplet-state natural organic matter loss by energy transfer and electron transfer pathways [J]. Environmental Science & Technology, 2013, 47(19):10987-10994.

[28] Cheng S, Zhang X, Yang X, et al. The multiple role of bromide ion in PPCPs degradation under UV/chlorine treatment [J]. Environmental Science & Technology, 2018,52(4):1806-1816.

[29] Zhou L, Deng H, Zhang W, et al. Photodegradation of sulfamethoxazole and photolysis active species in water under uv-vis light irradiation [J]. Fresenius Environmental Bulletin, 2015,24(5): 1685-1691.

[30] Chen Y, Jiang X, Xiao K, et al. Enhanced volatile fatty acids (VFAs) production in a thermophilic fermenter with stepwise pH increase– Investigation on dissolved organic matter transformation and microbial community shift [J]. Water Research, 2017,112:261-268.

[31] ter Laak T L, Gebbink W A, Tolls J. The effect of pH and ionic strength on the sorption of sulfachloropyridazine, tylosin, and oxytetracycline to soil [J]. Environmental Toxicology and Chemistry: An International Journal, 2006,25(4):904-911.

The indirect photodegradation of sulfapyridazine in seawater.

DOU Qi-wei, DUAN Jia-qi, TANG Xin-yu, YAO Qing-zhen, SU Rong-guo*

(College of Chemistry and Chemical Engineering, Ocean University of China, Qingdao 266100, China)., 2023,43(1)：190～196

The indirect photodegradation behavior and mechanism of SCP were studied by using CDOM that as the main photosensitizer. The effects of CDOM composition, salinity and pH on indirect photodegradation of SCP were analyzed. When CDOM concentration increased, the indirect photodegradation rate of SCP accelerated obviously. Upon irradiation, CDOM produced a variety of active substances, and the different active substances had different contribution rates to the indirect photodegradation of SCP, especially the3CDOM*played a major role in the indirect photodegradation of SCP and its contribution rate was up to 77.94%. The CDOM used in the experiment was composed of four components, including three terrestrial humus (C1, C2, C3) and the marine humus (C4). The order of the correlation between the SCP removal rates and the four fluorescence components was C3>C2>C4>C1. The correlation between four fluorescence components and [3CDOM*] were also ranked as the former. C3 and C2 components had a significant correlation with [3CDOM*] (2>0.97), that demonstrated the C3 and C2 components made great contributions to the production of [3CDOM*]. Salinity and pH value had significant effects on indirect photodegradation of SCP. When the salinity of the solution was equal to 15‰, the SCP indirect photodegradation rate reached a maximum. In the low salinity range (0～15‰), the promotion effect of ionic strength was greater than the inhibition effect of inorganic anions that made the indirect photodegradation rate greater with the increase of salinity. When the pH of the solution was equal to (5.00±0.10), the SCP indirect photodegradation rate reached a maximum. The indirect photodegradation rate of SCP slowed down with the increase of pH value of the solution, and the neutral and alkaline environments had negative effects on the indirect photodegradation of SCP.

sulfachloropyridazine；indirect photodegradation；CDOM composition；environmental factor

X703,X55

A

1000-6923(2023)01-0190-07

窦琦玮(1998-),女,山东潍坊人,中国海洋大学硕士研究生,主要研究方向为海洋污染生态化学.

2022-06-02

NSFC-山东联合基金资助项目(U1906210);国家重点研发计划项目(2016YFC1402101)

*责任作者, 教授, surongguo@ouc.edu.cn