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目录 contents

    摘要

    采用高级氧化(advanced oxidation processes, AOPs)技术去除水体中的头孢类抗生素。选取头孢类抗生素中的典型物质头孢氨苄(CFX)为研究对象,探讨了其在UV/过硫酸盐(UV/PS)体系中的降解特性。结果表明,pH=7.0时,UV/PS体系中 S O 4 - · 和·OH均对CFX有降解作用,且其与CFX反应的二阶速率常数分别为 9.8 ± 0.4 × 10 9 L ( m o l s ) - 1 1.05 ± 0.7 × 10 10 L ( m o l s ) - 1 。PS投加量的增加可加速CFX的降解和矿化,且酸性条件可促进CFX降解。水体基质Cl-的存在对CFX的降解起到了低浓度抑制高浓度促进的作用, H C O 3 - 和自然有机质(NOM)的存在对CFX的降解稍有抑制。在实际水样中的应用研究表明,UV/PS体系可以有效降解和矿化实际地表水样(SW)和实际废水样(WW)中的CFX,具有较好的应用前景。

    Abstract

    Advanced corresponding author:oxidation processes (AOPs) were applied to remove cephalosporin antibiotics in surface water (SW) and wastewater (WW). In this study, cefalexin (CFX), a typical cephalosporin antibiotic, was selected as the objective contaminant, and its degradation characterization by UV/persulfate (UV/PS) was investigated. Results showed that both S O 4 - · and ·OH contributed to CFX degradation at pH 7.0 and the corresponding second-order rate constants were 9.8 ± 0.4 × 10 9 L ( m o l s ) - 1 , 1.05 ± 0.7 × 10 10 L ( m o l s ) - 1 , respectively. The increase of PS dosage could accelerate the degradation and mineralization of CFX, and acidic condition could promote the CFX degradation. The effects of water matrices on CFX degradation were also investigated here. At low Cl- concentration, CFX degradation was inhibited, while high Cl- concentration could promote CFX degradation. Both H C O 3 - and natural organic matter (NOM) showed a slight inhibition on the degradation of CFX. Moreover, UV/PS could effectively degrade and mineralize the CFX in real SW and WW samples, implying its great potential in future.

    文章栏目:水污染防治

    李珂, 刘振鸿, 钱雅洁, 等. 基于硫酸根自由基的高级氧化对头孢氨苄的降解特性 [J]. 环境工程学报, 2019, 13(1): 40-48.

    我国是抗生素生产和使用大国,每年抗生素的使用量达21万t[1]。但由于我国对抗生素的生产、使用和处置缺乏有效的监管机制,导致排放到环境中的抗生素及其代谢产物不断增加,污染农作物土壤、地下水、地表水、甚至饮用[2]。环境中低浓度的抗生素暴露会诱导产生抗生素耐药菌及抗性基因,严重威胁生态安全及公众健[3]。头孢类抗生素是一种重要的抗生素,已广泛用于人类和兽类疾病和感染的治疗。但是由于传统的污水处理厂对抗生素等微污染物的去除率比较低,大部分的头孢类抗生素不能被污水处理系统有效去除,从而进入自然水体。近些年的环境调查发现,头孢类抗生素在污水处理厂出水、地表水、沿海水域等水体中普遍检[4]。与此同时,天然水体中的细菌长期暴露于痕量水平的抗生素中,引起了人们对抗生素诱导细菌群体产生抗菌抗性基因潜在风险的担忧。据世界卫生组织的统计,在我国,部分头孢类抗性基因的检出率高达70%[5]。因此,高效的处理技术去除水体中的头孢类抗生素是目前的研究热点。

    高级氧化是去除水体中抗生素类污染物的重要适用技术。近年来,基于UV激活过硫酸盐的新型高级氧化工艺受到广泛关注。相比于其他激活手段,如热激活、金属活化、臭氧活化、超声活化等,UV激活效率高、副产物少、反应条件更温[6]。同时,相比于·OH(E0 =1.9~2.7 V), S O 4 - · 具有更高的氧化性(E0 =2.5~3.1 V),且UV/PS具有更高的量子产率( Φ P S = 5.35 m o l ( k W h ) - 1 , Φ H 2 O 2 = 3.82 m o l ( k W h ) - 1 ( 1 E i n s t e i n = 0.130 8 k W h ) );相同光照条件下,UV/PS自由基产率更[7]。相对H2O2而言,PS的稳定性更高,反应过程中作用时间更长,适用的pH范围更广。众多的研[8,9,10,11,12,13,14]表明, S O 4 - · 可以降解内分泌干扰化合物、氯化化合物、药物、全氟化合物和藻毒素等,具有广阔的应用前景。

    S 2 O 8 2 - + h v 2 S O 4 - ε S 2 O 8 2 - 254 = 22.1 L c m 2 m o l - 1 Φ S 2 O 8 2 - = 5.35 m o l ( k W h ) - 1
    H 2 O 2 + h v 2 O H ε H 2 O 2 254 = 17.9 ~ 19.6 L c m 2 m o l - 1 Φ H 2 O 2 = 3.82 m o l ( k W h ) - 1

    本研究采用UV/PS体系降解水体中的CFX,确定CFX和 S O 4 - · 之间反应的二级速率常数,考察其降解动力学特性及不同水质条件对其降解的影响,旨在为水体中头孢抗生素的削减提供参考。

  • 1 材料与方法

    1
  • 1.1 实验原料

    1.1

    头孢氨苄(CFX, >98%)。过硫酸钠(PS, 99.9%)、双氧水(H2O2, 质量分数30%)、氯化钠(NaCl, 99%)、碳酸氢钠(NaHCO3, 99.7%)、磷酸氢二钠(Na2HPO4, 99.9%)、磷酸二氢钠(NaH2PO4, 99.9%)、硫代硫酸钠(Na2S2O3, 99.99%)、腐殖酸(humic acid, HA)、乙醇、叔丁醇等试剂均采用色谱纯。分析用甲醇、甲酸、乙酸为色谱级。实验溶液配置及反应用水均采用Milli-Q纯化系统(18.2 mΩ·cm)制备的超纯水。

  • 1.2 实验装置

    1.2

    实验过程中使用的仪器设备:pH计(FE28K, Mettler Toledo);高效液相色谱(SPD-20A, Shimadzu Corporation, Kyoto Japan)配二极管阵检测器(DAD)和Zorbax SB-C18色谱柱(4.6 mm×250 mm×5 μm);紫外-可见分光光度计(TU-1810, 北京普析通用仪器有限公司);TOC分析仪(multi N/C® 3100, analytikjena)。

  • 1.3 实验方法

    1.3

    CFX降解动力学实验在100 mL圆柱形石英反应器中进行,反应器配置1个4 W的低压紫外汞灯(输出波长为254 nm)及磁力搅拌器以维持反应溶液的均匀混合。反应过程中反应器所接受光强根据草酸铁钾(K3Fe(C2O4)3·3H2O)[15]进行测定,测得辐照强度为0.18×10-8 kWh·(L·s)-1(1 Einstein·(L·s)-1=0.130 8 kWh·(L·s)-1)。降解动力学实验使用10 mmol·L-1的磷酸盐缓冲液调节反应初始pH,分别将不同浓度的Cl- H C O 3 - 和NOM添加至反应体系中,探究水体基质对CFX降解的影响。实验过程中每隔30 ~180 s取样1 mL,取出的样品加入0.1 mL的硫代硫酸钠,以淬灭反应剩余的自由基,并在4 ℃下保存,48 h内用高效液相色谱仪进行定量分析。在实际水体CFX的降解实验中,分别选取实际地表水样(SW)和实际废水样(WW)作为背景溶液模拟实际水体,取出的水样在定量分析前通过0.45 μm滤膜,并在4 ℃下储存。所有实验重复3次。

    UV/PS体系的鉴定实验通过添加乙醇(EtOH, k O H / E t O H = 1.2 ~ 2.8 × 10 9 L ( m o l s ) - 1 , k S O 4 - · / A S = 4.9 × 10 9 L ( m o l s ) - 1 )[16,17]及叔丁醇(TBA, k O H / T B A = 3.8 ~ 7.6 × 10 8 L ( m o l s ) - 1 )[17]来鉴定体系中 S O 4 - · 和·OH的存在。

    速率常数的测定通常采用竞争实验来完成。由于UV/PS体系中含· O H S O 4 - · 2种主导自由基,无法通过竞争实验同时测定 k O H / C F X k S O 4 - · / C F X ,故本研究引入了以·OH为主的UV/H2O2体系。在UV/H2O2体系中,竞争实验使用硝基苯(NB, k O H / N B = 3.0 × 10 9 L ( m o l s ) - 1 )[16,18,19]作为·OH的竞争化合物,以测定 k O H / C F X 值。在UV/PS体系中,竞争实验使用苯甲醚(AS, k S O 4 - · / A S = 4.9 × 10 9 L ( m o l s ) - 1 )[18,20]作为 S O 4 - · 的竞争化合物,以测定 k S O 4 - · / C F X 值。在UV/H2O2和UV/PS体系中,CFX的降解由直接光解和间接光解共同作用,其中间接光解主要是由自由基氧化引起的,对于UV/H2O2和UV/PS体系,·OH和·OH、 S O 4 - · 分别是参与氧化的主导自由基。

    在UV/H2O2体系中,CFX的降解可以用式(1)表示:

    - d [ C F X ] d t = k o b s / C F X [ C F X ] = k d / C F X [ C F X ] + k i / C F X [ C F X ] = k d / C F X [ C F X ] + k O H / C F X [ O H ] [ C F X ]
    (1)

    NB作为反应中的竞争化合物,它的降解可以表示成相似的方程式:

    - d [ N B ] d t = k o b s / N B [ N B ] = k d / N B [ N B ] + k i / N B [ N B ] = k d / N B [ N B ] + k O H / N B [ O H ] [ N B ]
    (2)

    式中:kobs/CFXkobs/NB为CFX和NB的表观降解速率常数,s-1kd/CFXkd/NB分别为CFX和NB的直接光解速率常数,s-1ki/CFXki/NB为CFX和NB的间接光解速率常数,s-1k·OH/CFXk·OH/NB为CFX和NB及·OH的二级速率常数。

    结合式(1)和式(2),得到:

    l n C F X t C F X 0 - k d / C F X t = l n N B t N B 0 - k d / N B t k O H / C F X k O H / N B
    (3)

    由于CFX和NB的降解符合伪一级动力学规律,故k·OH/CFX可以用式(4)表示:

    k O H / C F X = k O H / N B k i / C F X k i / N B
    (4)

    对于UV/PS体系,首先向反应体系中加入200 mmol·L-1叔丁醇(TBA, k O H / T B A = 3.8 ~ 7.6 × 10 8 L ( m o l s ) - 1 )以淬灭反应体系中的·OH,然后使用类似的方法测定 S O 4 - · 与CFX反应的反应速率常数。 k S O 4 - · / C F X 可以用式(5)表示:

    k S O 4 - · / C F X = k S O 4 - · / A S k i / C F X k i / A S
    (5)
  • 1.4 分析方法

    1.4

    实验中PS浓度采用分光光度[21] 测定。在NaHCO3存在的条件下,PS与KI反应生成黄色络合物,通过紫外-可见分光光度计(TU-1810, 北京普析通用仪器有限公司)于400 nm波长处进行测定。

    实验中CFX浓度采用岛津高效液相色谱仪及C-18色谱柱进行定量分析。流动相为甲醇和0.1%的甲酸水溶液,V(甲醇)/V(0.1%甲酸水)=25/75,流速1.0 mL·min-1,柱温35 ℃,检测波长254 nm,保留时间12 min。

  • 2 结果与讨论

    2
  • 2.1 CFX的降解速率及矿化率

    2.1

    pH为7.0、CFX初始浓度0.1 mmol·L-1条件下,检测CFX的自然水解速率、PS单独氧化速率、单独UV光解速率以及在UV/PS体系中不同PS投加量下的降解速率和矿化率,结果如图1所示。由图1(a)可以看出,体系中CFX的降解与反应时间呈良好的线性关系(R2 >0.95)。且经测定,反应过程中PS浓度变化不大,故本实验可以通过伪一级动力学方程计算CFX降解的表观速率常数:

    - d C F X / d t = k o b s C F X
    (6)

    式中:[CFX]为CFX的摩尔浓度,mmol·L-1kobs为伪一级动力学常数;t为反应时间,h。

    实验将CFX自然水解效果、PS单独氧化效果、单独UV光解效果与UV/PS体系降解效果进行了对比,发现自然水解和PS单独氧化对CFX的降解速率很小,远小于CFX的直接光解速率,可忽略不计。在此基础上,实验通过向体系中投加不同浓度的PS,来探究PS投加量对体系中CFX降解和矿化的影响。由图1(a)可以看出,与单独UV光解相比,随着初始PS浓度从0 mmol·L-1到2 mmol·L-1的增加,体系中CFX的降解速率有显著的提升。而PS直接氧化对CFX的降解速率几乎无影响,因此, S O 4 - · 的形成随着初始PS浓度的增加而增加,从而加快了CFX的降解。由图1(b)可以看出,单独UV对CFX的矿化率很低,难以实现彻底矿化。而PS的加入可以有效提高矿化效率,当PS投加量为10 mmol·L-1时8 min内可以实现CFX的完全矿化,对于实际水样中CFX的削减具有重要意义。

    图1
                            UV/PS体系中CFX的降解与矿化

    图1 UV/PS体系中CFX的降解与矿化

    Fig. 1 CFX degradation and mineralization by UV/PS

  • 2.2 体系中自由基的鉴定及速率常数的测定

    2.2

    在UV/PS体系中,PS被激活可以产生 S O 4 - · S O 4 - · 与水反应或者在碱性条件下与OH-反应时可以产生·OH,见反应式(7)和式(8)。本研究通过投加EtOH及TBA对体系中主要自由基进行了鉴定,结果见图2

    S O 4 - + H 2 O O H + H S O 4 -
    (7)
    S O 4 - + O H - O H + S O 4 2 -
    (8)

    由图2可以看出,UV/PS体系中EtOH和TBA的存在均不同程度上抑制了CFX的降解,且EtOH对CFX的抑制能力强于TBA。由于EtOH可以同时淬灭体系中的·OH及 S O 4 - · ,而TBA只可以淬灭·OH,因此,二者的存在所导致体系中CFX降解速率的差异表明体系中 S O 4 - · 和·OH是共同存在的。同时,从CFX降解的kobs可以看出,添加TBA和EtOH后,CFX降解的kobs逐级降低。添加TBA与UV/PS之间的差值即可表示体系中·OH的作用,而添加EtOH与UV/PS之间的差值即可表示体系中 S O 4 - · 和·OH的共同作用,从kobs递减的差值可以判断 S O 4 - 在降解CFX的过程中发挥了重要作用。

    图2
                            TBA和EtOH对CFX降解的影响及CFX降解的kobs

    图2 TBA和EtOH对CFX降解的影响及CFX降解的kobs

    Fig. 2 Effects of TBA and EtOH on CFX degradation and the kobs

    为了进一步探究CFX与 S O 4 - 和·OH反应的动力学特性,本研究采用竞争动力学方[22]测定了CFX的二级反应速率常数 k O H / C F X k S O 4 - / C F X 的值。目前该方法广泛应用于测算很多物质与自由基的二级反应速率常数。实验使用硝基苯(NB, k O H / N B = 3.0 × 10 9 L ( m o l s ) - 1 )和苯甲醚(AS, k S O 4 - / A S = 4.9 × 10 9 L ( m o l s ) - 1 )作为参照物质,通过CFX的竞争实验,得到UV/H2O2体系中 k i / C F X = 7.0 ± 0.1 × 10 - 4 s - 1 k i / N B = 2.0 ± 0.1 × 10 - 4 s - 1 ,UV/PS体系中 k i / C F X = 8.0 ± 0.1 × 10 - 4 s - 1 k i / A S = 4.0 ± 0.1 × 10 - 4 s - 1 。在此基础上,结合1.3.3的计算方法,计算得到UV/H2O2和UV/PS体系中CFX与·OH 和 S O 4 - 自由基的二阶反应速率常数,分别为 k O H / C F X = 1.05 ± 0.7 × 10 10 L ( m o l s ) - 1 k S O 4 - / C F X = 9.8 ± 0.4 × 10 9 L ( m o l s ) - 1 。TOTH[23]通过线性加速器(LINAC)电子脉冲辐射分解系统,计算得出基于硫酸根的高级氧化体系中CFX与 S O 4 - 的二阶反应速率常数 k S O 4 - / C F X = 2.4 × 10 9 L ( m o l s ) - 1 ,此结论与本研究所得结果相近。

  • 2.3 pH对CFX降解的影响

    2.3
    图3
                            不同初始pH下CFX降解的kobs

    图3 不同初始pH下CFX降解的kobs

    Fig. 3 kobs of CFX degradation under different initial pH

    3是不同pH条件下CFX在UV/PS体系中的降解速率。由图3可以看出,pH=3时CFX降解的kobs最高,达到(6.80±0.22) h-1。且随着溶液pH的升高,CFX降解的kobs逐渐降低。较低pH下CFX降解较快,原因可能是在酸性条件下,CFX以质子态存在(CFX0⇌CFX-+H+pKa=6.88)[24],相对碱性条件下带负电CFX-,CFX0更易于和带负电的 S O 4 - 自由基反应。此外,酸性条件下,PS可以发生酸催化反[25],见反应式(9)和式(10),PS的解离增加,增大了体系中 S O 4 - 的产生量,从而促进了CFX的降解。此现象与 S O 4 - 降解氟喹诺酮抗生[26]和三氯乙烷的研究结[27]相近。

    随着pH的升高,溶液中OH-的浓度也随之增加,溶液中的OH-会与 S O 4 - 反应生成·OH,见反应式(11)。相对于 S O 4 - 来说,·OH不具有选择性,倾向于脱氢及加成反应,而 S O 4 - 倾向于电子转移。因此,当溶液中 S O 4 - 的浓度降低而·OH浓度增加时,中间产物更多地参与了自由基的争[26],从而抑制了CFX的降解。LIANG[28]在研究pH如何影响过硫酸盐对TCE的降解中也有类似的发现。

    S 2 O 8 2 - + H + H S 2 O 8 -
    (9)
    H S 2 O 8 - S O 4 - · + S O 4 2 - + H +
    (10)
    S O 4 - · + O H - O H + S O 4 2 -
    (11)
  • 2.4 水体基质的影响

    2.4

    自然水基质中含有众多的自然有机质(NOM)以及无机离子,如Cl- H C O 3 - 等,这些物质的存在可以在一定程度上影响CFX在基于 S O 4 - 的AOPs中的降解。本研究在相同实验条件下([CFX] = 0.1 mmol·L-1, [PS]=1 mmol·L-1, pH=7),通过向体系中投加不同浓度的Cl- 、HCO3- 和腐殖酸(HA)来探究水体基质对CFX降解的影响。

    如图4所示,不同的Cl-浓度对CFX降解的影响不同。Cl-浓度较低时对CFX的降解有一定的抑制作用,而当Cl-浓度达到10 mmol·L-1时,促进了CFX的降解。Cl-浓度较低时抑制了CFX的降解,可能是当溶液中存在少量的Cl-时, S O 4 - ·较高的氧化还原电位(E0 =2.60 V)会将Cl-氧化成·Cl(E0 =2.41 V)10],见反应式(12)。·Cl的氧化还原电位较低,从而对体系中的 S O 4 - · 起到了竞争作用,故降低了 S O 4 - · 对CFX的去除速率。

    图4
                            不同浓度水体基质下CFX降解的kobs

    图4 不同浓度水体基质下CFX降解的kobs

    Fig. 4 kobs of CFX degradation under different concentration of water matrices

    S O 4 - · + C l - C l + S O 4 2 - k = 2.7 × 10 8 L ( m o l s ) - 1
    (12)

    随着体系中Cl-浓度的增大,溶液中的·Cl发生链式反应,会生成部分衍生自由基,如 C l 2 - ·、ClOH-·等。相对·OH、 S O 4 - · 来说,这些衍生自由基虽然氧化性稍弱,但对污染物的作用具有很强的选择性且其浓度远远大于 S O 4 - · 。据此分析,通过一系列链式反应产生的衍生自由基与头孢抗生素分子结构中含有多个富电子的 N、O、S 基团(如CFX分子侧链上的伯胺)反应,促进了CFX的降解。同样的现象在WANG[29]关于偶氮染料在高浓度Cl-(> 50 mmol·L-1)存在时降解增强的研究中也有出现。

    HCO3-和HA的存在抑制了CFX的降解,且抑制作用随着浓度的增加而增强。 H C O 3 - S O 4 - · 和·OH 反应的速率常数分别为 1.6 × 10 6 L ( m o l s ) - 1 8.5 × 10 6 L ( m o l s ) - 1 [16],因此,会与CFX竞争与自由基的反应,从而使CFX的降解受到抑制。而HA在254 nm处,对光有一定的吸收,因此,HA作为一种内部过滤作用影响了光解过程,从而降低了PS活化生成 S O 4 - · 的效率。此外,NOM同样也存在着与CFX竞争自由基的效应,其与 S O 4 - · 和·OH速率常数分别为 k S O 4 - / N O M = 2.4 × 10 7 L ( m o l s ) - 1 [7] k O H / N O M = 3.0 × 10 8 L ( m o l s ) - 1 [16],见反应式(13)和式(14)。因此, H C O 3 - 和HA的存在均可降低体系中自由基的浓度,导致CFX降解的kobs降低。

    S O 4 - · + N O M k = 2.4 × 10 7 L ( m o l s ) - 1
    (13)
    O H + N O M k = 3.0 × 10 8 L ( m o l s ) - 1
    (14)
  • 2.5 实际水体中的应用

    2.5

    研究采用实际废水样(WW)和实际地表水样(SW)来模拟实际水体中UV/PS体系对CFX的降解及矿化效果。WW和SW样品分别取自上海市某污水处理厂二沉池出水和太湖湖水,其基本水质参数如表1所示,且在这2种实际水样中没有检测到CFX。

    在相同实验条件下([CFX] = 0.1 mmol·L-1,[PS]= 1 mmol·L-1, pH=7),UV/PS体系对CFX的降解如图5(a)所示。1 200 s内,该体系对WW和SW中CFX的去除率分别达到了87%和96%,且kobs/SW > kobs/WW。其主要是因为WW中存在大量的水体基质,一方面吸收了紫外光辐射,降低了溶液中生成的活性自由基总浓度;另一方面水体基质会与活性自由基发生反应,与CFX竞争自由基,降低了与CFX反应的自由基浓度,进而导致kobs/WW降低。

    表1 实际水样的组分

    Table 1 Characteristics of the water samples

    实际水样种类pHDOC/ (mg·L-1)Cl-/ (mmol·L-1) N O 3 - / (mmol·L-1) P O 4 3 - / (mmol·L-1)
    SW7.152.380.120.050.04
    WW7.0415.691.330.97< 0.02

    在UV/PS体系对WW和SW中CFX的矿化的研究中,实验加大了PS投加量并延长了反应时间([CFX] = 0.1 mmol·L-1, [PS]= 10 mmol·L-1, pH=7),矿化效果见图5(b)。研究表明,当PS投加量为10 mmol·L-1时,100 min内0.1 mmol·L-1 CFX的矿化率可达100%,该结论可以为实际废水和实际地表水中CFX的削减提供理论依据。在此基础上,从降解成本角度进一步评估了UV/PS的实际应用性。首先相对于热激活、金属活化、臭氧活化、超声活化等激活方式,UV激活效率[6]。其次,由于生产单位质量PS的能耗与生产单位质量H2O2的能耗基本一致,分别是10.81 kWh·kg-1和10.80 kWh·kg-1[30],而相对于UV/H2O2体系而言,UV/PS体系具有更高的量子产率( Φ P S = 5.35 m o l ( k W h ) - 1 , Φ H 2 O 2 = 3.82 m o l ( k W h ) - 1 ),相同光照条件下,UV/PS自由基产率更大,效率更[7]。因此,UV/PS技术在实际水体CFX的削减控制中具有较高的应用前景。

    图5
                            SW和WW中CFX的降解和矿化

    图5 SW和WW中CFX的降解和矿化

    Fig. 5 CFX degradation and mineralization in SW(surface water)and WW(wastewater)samples

  • 3 结论

    3

    1) UV/PS体系可以有效去除实际水样中微量的头孢类抗生素CFX,且kobs/SW > kobs/WW。此外,PS投加量为10 mmol·L-1时,UV/PS体系可在100 min内实现0.1 mmol·L-1 CFX的完全矿化。

    2) pH为7.0时,UV/PS体系中 S O 4 - · 和·OH同时存在并降解CFX,CFX与 S O 4 - · 和·OH的二阶反应速率常数分别为 k S O 4 - / C F X = 9.8 ± 0.4 × 10 9 L ( m o l s ) - 1 k O H / C F X = 1.05 ± 0.7 × 10 10 L ( m o l s ) - 1 ,但该体系中 S O 4 - · 对CFX降解的贡献大于·OH。

    3) PS投加量的增加可加速CFX的降解和矿化,且酸性条件更有利于CFX的降解。

    4) 水体基质Cl-的存在对CFX的降解起到了低浓度抑制高浓度促进的作用。NOM和HCO3-是天然水体中主要的自由基清除剂,可以同时消耗 S O 4 - · 和·OH自由基,而且NOM对光有一定的吸收,因而二者的存在抑制了CFX的降解。

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      ZHU Y G, JOHNSON T A, SU J Q, et al. Diverse and abundant antibiotic resistance genes in Chinese swine farms[J]. Pans, 2013, 110(9): 3435-3440.

    • 2

      HAO R, ZHAO R, QIU S, et al. Antibiotics crisis in China[J]. Science, 2015, 348(6239): 1100-1101.

    • 3

      ZHANG Q, JIA A, WAN Y, et al. Occurrences of three classes of antibiotics in a natural river basin: Association with antibiotic-resistant Escherichia coli[J]. Environmental Science & Technology, 2014, 48(24): 14317-14325.

    • 4

      CHEN J, SUN P, ZHANG Y, et al. Multiple roles of Cu(II) in catalyzing hydrolysis and oxidation of β-Lactam antibiotics[J]. Environmental Science & Technology, 2016, 50(22): 12156-12165.

    • 5

      ORGANIZATION W H. Antimicrobial resistance: Global report on surveillance[J]. Australasian Medical Journal, 2014, 7(4):238-239.

    • 6

      郭洪光, 刘杨, 张永丽. 紫外激活过硫酸钠降解环境雌激素17β-雌二醇分析[J]. 东北大学学报(自然科学版), 2016, 37(6): 880-885.

    • 7

      NETA P, HUIE R E, ROSS A B. Rate constants for reactions of inorganic radicals in aqueous solution[J]. Journal of Physical & Chemical Reference Data, 1988, 17(3): 1027-1284.

    • 8

      TIM K L, WEI C, NIGEL J D G. The aqueous degradation of butylated hydroxyanisole by UV/S2 O 8 2 - : Study of reaction mechanisms via dimerization and mineralization[J]. Environmental Science & Technology, 2007, 41(2): 613-619.

    • 9

      GUAN Y H, MA J, LI X C, et al. Influence of pH on the formation of sulfate and hydroxyl radicals in the UV/peroxymonosulfate system[J]. Environmental Science & Technology, 2011, 45(21): 9308-9314.

    • 10

      RACHEL H W, PAUL G T, RICHAED L J, et al. Oxidation of chlorinated ethenes by heat-activated persulfate: Kinetics and products[J]. Environmental Science & Technology, 2007, 41(3): 1010-1015.

    • 11

      GUAN Y H, MA J, REN Y M, et al. Efficient degradation of atrazine by magnetic porous copper ferrite catalyzed peroxymonosulfate oxidation via the formation of hydroxyl and sulfate radicals[J]. Water Research, 2013, 47(14): 5431-5438.

    • 12

      GAO Y, GAO N, DENG Y, et al. Ultraviolet (UV) light-activated persulfate oxidation of sulfamethazine in water[J]. Chemical Engineering Journal, 2012, 195-196: 248-253.

    • 13

      HORI H, YAMAMOTO A, HAYAKAWA E, et al. Efficient decomposition of environmentally persistent perfluorocarboxylic acids by use of persulfate as a photochemical oxidant[J]. Environmental Science & Technology, 2005, 39(7): 2383-2388.

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      ANTONIOU M G, CRUZ A A D L, DIONYSIOU D D. Intermediates and reaction pathways from the degradation of microcystin-LR with sulfate radicals[J]. Environmental Science & Technology, 2010, 44(19): 7238-7244.

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      KUHN H J, BRASLAVSKY S E, SCHMIDT R. Chemical actinometry[J]. Pure & Applied Chemistry, 1989, 61(2): 187-210.

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      BUXTON G V, GREENSTOCK C L, HELMAN W P, et al. Critical review of rate constants for reactions of hydrated electrons, hydrogen atoms and hydroxyl radicals (⋅OH/⋅O-) in aqueous solution[J]. Journal of Physical & Chemical Reference Data, 1988, 17(2): 513-886.

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      ANIPSITAKIS G P, DIONYSIOU D D. Radical generation by the interaction of transition metals with common oxidants[J]. Environmental Science & Technology, 2004, 38(13): 3705-3712.

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      NETA P, MADHAVAN V, ZEMEL H, et al. Rate constants and mechanism of reaction of S O 4 - ·with aromatic compounds[J]. Journal of the American Chemical Society, 1977, 8(14): 163-164.

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      O'NEILL P, STEENKEN S, SCHULTE-FROHLINDE D. Formation of radical cations of methoxylated benzenes by reaction with OH radicals, Tl2+, Ag2+, and S O 4 - · in aqueous solution. Optical and conductometric pulse radiolysis and in situ radiolysis electron spin resonance study[J]. Journal of Physical Chemistry, 1975, 79(25): 2773-2779.

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      LIANG C, HUANG C F, MOHANTY N, et al. A rapid spectrophotometric determination of persulfate anion in ISCO[J]. Chemosphere, 2008, 73(9): 1540-1543.

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      ZHANG R, SUN P, BOYER T H, et al. Degradation of pharmaceuticals and metabolite in synthetic human urine by UV, UV/H2O2 and UV/PDS[J]. Environmental Science & Technology, 2015, 49(5): 3056-3066.

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      TOTH J E, RICKMAN K A, VENTER A R, et al. Reaction kinetics and efficiencies for the hydroxyl and sulfate radical based oxidation of artificial sweeteners in water[J]. Journal of Physical Chemistry A, 2012, 116(40): 9819-9824.

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      CHEN J, FANG C, XIA W, et al. Selective transformation of β-lactam antibiotics by peroxymonosulfate: Reaction kinetics and nonradical mechanism[J]. Environmental Science & Technology, 2018, 52(3): 1461-1470.

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      HOUSE D A. Kinetics and mechanism of oxidations by peroxydisulfate[J]. Chemical Reviews, 1961, 62(3): 185-203.

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      郭洪光, 高乃云, 张永丽, 等. 热激活过硫酸盐降解水中典型氟喹诺酮抗生素分析[J]. 四川大学学报(工程科学版), 2015, 47(2): 191-197.

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      LIANG C J, BRUELL C J, MARLEY M C, et al. Thermally activated persulfate oxidation of trichloroethylene (TCE) and 1,1,1-trichloroethane (TCA) in aqueous systems and soil slurries[J]. Journal of Soil Contamination, 2003, 12(2): 207-228.

    • 28

      LIANG C, WANG Z S, BRUELL C J. Influence of pH on persulfate oxidation of TCE at ambient temperatures[J]. Chemosphere, 2007, 66(1): 106-113.

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      WANG Z, YUAN R, GUO Y, et al. Effects of chloride ions on bleaching of azo dyes by Co2+/oxone regent: Kinetic analysis[J]. Journal of Hazardous Materials, 2011, 190(1/2/3): 1083-1087.

    • 30

      QIAN Y, GUO X, ZHANG Y, et al. Perfluorooctanoic acid degradation using UV-persulfate process: Modeling of the degradation and chlorate formation[J]. Environmental Science & Technology, 2016, 50(2): 772-781.

李珂

机 构:东华大学环境科学与工程学院,上海 201620

Affiliation:College of Environmental Science and Engineering, Donghua University, Shanghai 201620, China

角 色:第一作者

Role:First author

邮 箱:lkdhu8305@163.com

第一作者简介:李珂(1993— ),女,硕士研究生。研究方向:水污染控制理论与技术。E-mail:lkdhu8305@163.com

刘振鸿

机 构:东华大学环境科学与工程学院,上海 201620

Affiliation:College of Environmental Science and Engineering, Donghua University, Shanghai 201620, China

钱雅洁

机 构:东华大学环境科学与工程学院,上海 201620

Affiliation:College of Environmental Science and Engineering, Donghua University, Shanghai 201620, China

角 色:通讯作者

Role:Corresponding author

邮 箱:yqian@dhu.edu.cnyqian@dhu.edu.cn

作者简介:钱雅洁(1987— ),女,博士,讲师。研究方向:高级氧化处理技术等。E-mail:yqian@dhu.edu.cn

薛罡

机 构:东华大学环境科学与工程学院,上海 201620

Affiliation:College of Environmental Science and Engineering, Donghua University, Shanghai 201620, China

王麒

机 构:东华大学环境科学与工程学院,上海 201620

Affiliation:College of Environmental Science and Engineering, Donghua University, Shanghai 201620, China

李前

机 构:东华大学环境科学与工程学院,上海 201620

Affiliation:College of Environmental Science and Engineering, Donghua University, Shanghai 201620, China

金曙光,郑晓梅,张利田

角 色:中文编辑

Role:Editor

/html/teepc_cn/201808012/alternativeImage/8333ac24-77cd-45e7-b305-3428f07d56e2-F001.jpg
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实际水样种类pHDOC/ (mg·L-1)Cl-/ (mmol·L-1) N O 3 - / (mmol·L-1) P O 4 3 - / (mmol·L-1)
SW7.152.380.120.050.04
WW7.0415.691.330.97< 0.02
/html/teepc_cn/201808012/alternativeImage/8333ac24-77cd-45e7-b305-3428f07d56e2-F005.jpg

图1 UV/PS体系中CFX的降解与矿化

Fig. 1 CFX degradation and mineralization by UV/PS

图2 TBA和EtOH对CFX降解的影响及CFX降解的kobs

Fig. 2 Effects of TBA and EtOH on CFX degradation and the kobs

图3 不同初始pH下CFX降解的kobs

Fig. 3 kobs of CFX degradation under different initial pH

图4 不同浓度水体基质下CFX降解的kobs

Fig. 4 kobs of CFX degradation under different concentration of water matrices

表1 实际水样的组分

Table 1 Characteristics of the water samples

图5 SW和WW中CFX的降解和矿化

Fig. 5 CFX degradation and mineralization in SW(surface water)and WW(wastewater)samples

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  • 参 考 文 献

    • 1

      ZHU Y G, JOHNSON T A, SU J Q, et al. Diverse and abundant antibiotic resistance genes in Chinese swine farms[J]. Pans, 2013, 110(9): 3435-3440.

    • 2

      HAO R, ZHAO R, QIU S, et al. Antibiotics crisis in China[J]. Science, 2015, 348(6239): 1100-1101.

    • 3

      ZHANG Q, JIA A, WAN Y, et al. Occurrences of three classes of antibiotics in a natural river basin: Association with antibiotic-resistant Escherichia coli[J]. Environmental Science & Technology, 2014, 48(24): 14317-14325.

    • 4

      CHEN J, SUN P, ZHANG Y, et al. Multiple roles of Cu(II) in catalyzing hydrolysis and oxidation of β-Lactam antibiotics[J]. Environmental Science & Technology, 2016, 50(22): 12156-12165.

    • 5

      ORGANIZATION W H. Antimicrobial resistance: Global report on surveillance[J]. Australasian Medical Journal, 2014, 7(4):238-239.

    • 6

      郭洪光, 刘杨, 张永丽. 紫外激活过硫酸钠降解环境雌激素17β-雌二醇分析[J]. 东北大学学报(自然科学版), 2016, 37(6): 880-885.

    • 7

      NETA P, HUIE R E, ROSS A B. Rate constants for reactions of inorganic radicals in aqueous solution[J]. Journal of Physical & Chemical Reference Data, 1988, 17(3): 1027-1284.

    • 8

      TIM K L, WEI C, NIGEL J D G. The aqueous degradation of butylated hydroxyanisole by UV/S2 O 8 2 - : Study of reaction mechanisms via dimerization and mineralization[J]. Environmental Science & Technology, 2007, 41(2): 613-619.

    • 9

      GUAN Y H, MA J, LI X C, et al. Influence of pH on the formation of sulfate and hydroxyl radicals in the UV/peroxymonosulfate system[J]. Environmental Science & Technology, 2011, 45(21): 9308-9314.

    • 10

      RACHEL H W, PAUL G T, RICHAED L J, et al. Oxidation of chlorinated ethenes by heat-activated persulfate: Kinetics and products[J]. Environmental Science & Technology, 2007, 41(3): 1010-1015.

    • 11

      GUAN Y H, MA J, REN Y M, et al. Efficient degradation of atrazine by magnetic porous copper ferrite catalyzed peroxymonosulfate oxidation via the formation of hydroxyl and sulfate radicals[J]. Water Research, 2013, 47(14): 5431-5438.

    • 12

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