环境科学  2023, Vol. 44 Issue (7): 3957-3969   PDF    
引用本文
葛林科, 王子宇, 曹胜凯, 车晓佳, 朱超, 张蓬, 马宏瑞. 新污染物多环芳烃衍生物的来源、分布与光化学行为[J]. 环境科学, 2023, 44(7): 3957-3969.
GE Lin-ke, WANG Zi-yu, CAO Sheng-kai, CHE Xiao-jia, ZHU Chao, ZHANG Peng, MA Hong-rui. Critical Review on Environmental Occurrence and Photochemical Behavior of Substituted Polycyclic Aromatic Hydrocarbons[J]. Environmental Science, 2023, 44(7): 3957-3969.
新污染物多环芳烃衍生物的来源、分布与光化学行为
葛林科, 王子宇, 曹胜凯, 车晓佳, 朱超, 张蓬, 马宏瑞     
陕西科技大学环境科学与工程学院, 西安 710021
收稿日期: 2022-07-27; 修订日期: 2022-09-24
基金项目: 国家环境保护近岸海域生态环境重点实验室开放基金项目(202102); 国家自然科学基金项目(21976045, 22076112); 中国科学院海岸带环境过程与生态修复重点实验室开放基金项目(2020KFJJ03); 国家水体污染控制与治理科技重大专项(2017ZX07602-001)
作者简介: 葛林科(1980~), 男, 博士, 教授, 主要研究方向为新型有机污染物的环境光化学, E-mail: gelinke@sust.edu.cn
通信作者: 张蓬，E-mail: zhangpeng4477@sust.edu.cn
摘要: 环境中多环芳烃衍生物(SPAHs)来源广泛, 具有“三致”效应, 毒性与PAHs母体相当甚至高于母体, 是一类具有较高风险的新型有机污染物.大气和表层水体等环境介质中, 光化学降解是其重要的消减方式.在梳理文献的基础上, 发现SPAHs在来源、分布、行为和风险研究方面与传统污染物PAHs相比有了新的突破.通过总结新污染物SPAHs的环境分布特征与光化学行为研究的最新进展, 介绍了3类SPAHs的来源和形成机制, 重点评述了不同环境介质中的存在状况与分布特征, 讨论了水、冰等介质中SPAHs光化学转化动力学、反应路径以及影响因素, 最后对SPAHs的环境行为和风险研究进行了展望.
关键词: 多环芳烃衍生物      生成机制      分布特征      光化学行为      影响因素     
Critical Review on Environmental Occurrence and Photochemical Behavior of Substituted Polycyclic Aromatic Hydrocarbons
GE Lin-ke , WANG Zi-yu , CAO Sheng-kai , CHE Xiao-jia , ZHU Chao , ZHANG Peng , MA Hong-rui     
School of Environmental Science and Engineering, Shaanxi University of Science & Technology, Xi'an 710021, China
Abstract: Substituted polycyclic aromatic hydrocarbons (SPAHs) are a type of emerging pollutant that widely exist in the environment, which also exhibit carcinogenicity, mutagenicity, and teratogenicity. These pollutants belong to toxic pollutants because of their similar structures to polycyclic aromatic hydrocarbons (PAHs). Their environmental behavior and ecological risk have attracted increasing attention. Based on a literature review, we found a new breakthrough in the source, distribution, behavior, and risk of SPAHs with comparison to traditional pollutants PAHs. This paper reviewed the current research progress on the environmental occurrence and photochemical behavior of SPAHs. Their sources, formation mechanisms, and distribution characteristics in the multimedia environment were highlighted, and the photochemical transformation kinetics, pathways, and affecting factors of SPAHs in water, ice, and other media were discussed. Furthermore, the research prospects about the environmental behavior and risk of SPAHs were proposed.
Key words: substituted polycyclic aromatic hydrocarbons      formation mechanisms      distribution characteristics      photochemical behavior      affecting factors     

多环芳烃衍生物(substituted polycyclic aromatic hydrocarbons, SPAHs)是目前公认的新型有机污染物, 包括羟基多环芳烃(OH-PAHs)、硝基多环芳烃(nitro-PAHs)和氯代多环芳烃(Cl-PAHs)等(图 1).对于SPAHs的毒性、来源和分布等, 已有研究进行了总结[1~3].SPAHs具有致癌、致畸和致突变效应, 其毒性与PAHs母体相当, 甚至高于母体[4].例如, 相对于母体, OH-PAHs的细胞毒性更大[5], nitro-PAHs的致突变能力更强[1, 3], Cl-PAHs具有极高的致癌、致畸作用[2, 6].甚至, 某些SPAHs表现为光致毒性, 例如9-羟基芴(9-OHFl)[5]和1-硝基芘[7].并且SPAHs具有生物富集性, 人体可通过呼吸、饮食等多种途径摄入[8].因此, SPAHs被视为一类新型的高风险有机污染物, 对生态环境和人类健康具有一定的潜在威胁.

图 1 自然界中常见的3类多环芳烃衍生物(2~4环)的化学结构式 Fig. 1 Chemical structures of three classes of polycyclic aromatic hydrocarbon derivatives (2-4 rings) commonly found in natural media

环境中SPAHs类新型有机污染物普遍存在, 其在大气、土壤、水、沉积物和生物体内均有检出.并且, SPAHs容易在各介质之间发生迁移, 也可在水体和大气等介质中发生转化, 如光化学转化.光化学转化是污染物的一种重要环境行为, 环境中普遍存在着有机污染物吸收日光而发生的直接光解, 以及光敏剂存在下的间接光解[9].SPAHs在有机溶剂[9~12]、水[7]、固体表面[10, 12]和大气中的光化学行为已有报道, 并且冰雪环境中该类污染物的光化学行为也有研究, 例如冰中OH-PAHs的光降解及光修饰毒性, 以及纯水雪中nitro-PAHs的光降解动力学[13].以上研究表明, 光降解是SPAHs污染物重要的转化途径, 强烈影响其环境归趋和风险.

近年来, SPAHs的环境存在与光化学行为研究不断丰富和创新, 在其来源、分析方法、分布特征和光化学转化等方面取得了一些重要进展.对此, 前人仅总结了其中1~2类SPAHs在水或大气中(主要是大气中)的存在状况[14, 15], 以及nitro-PAHs的光化学行为[1, 3].鉴于此, 本文将梳理OH-PAHs、nitro-PAHs、Cl-PAHs的环境存在状况和光化学行为的最新研究进展, 全面评述大气、水、沉积物、土壤和生物体等5种介质中这3类SPAHs的存在状况与分布特征, 并着重讨论其在主要介质水和冰中的光化学转化机制与途径, 以及环境因子对SPAHs光化学行为的影响.

1 SPAHs来源和生成机制

自然界中SPAHs的种类繁多, 来源广泛, 存在着很大的“储存库”.除人为排放(不完全燃烧和氯化副产物等)[16], SPAHs还存在着广泛的“天然来源”, 既可以通过PAHs的化学转化(热反应)、光化学转化或生物化学作用生成, 也可通过生物代谢由PAHs转化生成, 并随排泄物进入环境[17].

大多OH-PAHs的产生机制是来自PAHs的光化学反应和生物转化.水体、冰相和大气环境中PAHs能发生光催化或与·OH反应生成OH-PAHs[13]; 土壤、沉积物和生物体中一些细菌、真菌和微生物在自身的氧化酶作用下也会把PAHs转化为OH-PAHs[18, 19]. OH-PAHs还是PAHs经多种途径进入人体经肝脏代谢后的产物, 并以结合态的葡萄醛酸盐和磺酸盐等形式存在, 最后经尿液或粪便排出体外[20].目前, OH-PAHs被广泛作为生物标记物以监测人体是否有PAHs暴露风险, 用于评估环境和人类健康风险.郑金平等[21]将焦炉工人尿液中1-羟基芘作为PAHs内暴露标志物, 评估了机体长期暴露于PAHs对炎症反应和肺功能的影响.此外, Kuang等[22]还通过尿液中1-羟基芘含量, 评估了焦炉工人暴露于PAHs对DNA氧化损伤的影响.

nitro-PAHs主要由化石燃料直接燃烧排放, 以及通过大气中PAHs的光化学反应生成, 不同的反应途径会产生不同的nitro-PAHs异构体.nitro-PAHs与PAHs的形成过程类似, 两者均可在化石燃料不充分燃烧的情况下生成, 燃烧过程中高温引起PAHs发生亲电硝化反应, 生成nitro-PAHs[17].Wada等[23]和Miller-Schulze等[24]分别从汽油及柴油车尾气中检测出了nitro-PAHs, 汽车尾气中含量最高的nitro-PAHs是1-硝基芘[25].在中国东北地区, 冬季供暖燃煤也为大气中nitro-PAHs的环境浓度做出了显著贡献[26].除汽车燃油燃烧排放外, 部分nitro-PAHs通过与大气中自由基发生光化学反应生成, 根据自由基的不同, 可分为两种反应类型(图 2).①日间反应：因为白天大气中NO2和·OH浓度很高, 在光照射下, ·OH加成到PAHs上, 继而与NO2发生反应脱去一个水分子, 生成nitro-PAHs; ②夜间反应：傍晚之后N2O5会分解成硝基自由基和1O2, 加成到PAHs上, 再与NO2反应之后脱去一个硝酸分子, 生成nitro-PAHs.

改自文献[27] 图 2 SPAHs由母体PAHs生成的转化过程 Fig. 2 Transformation process of the SPAHs formed from the parent PAHs

Cl-PAHs是在人为的多种燃烧过程中形成和释放的副产物, 其主要释放源有固体废物焚烧[28]、一次和二次冶金处理[16]、水泥窑生产协同处理固体废物过程[15], 以及汽车尾气的排放[29], 其生成机制见图 3.Eklund等[30]报道了城市生活垃圾焚烧炉和燃煤锅炉产生的烟气中Cl-PAHs质量浓度在0.01~23 μg·m-3范围内.Wang等[31, 32]指出聚氯乙烯(PVC)在燃烧过程中会产生Cl-PAHs, 萘(Nap)、荧蒽(Flu)、菲(Phe)、蒽(Ant)、芘(Pyr)和芴(Flu)等母体PAHs与聚氯乙烯(PVC)燃烧释放出的HCl气体反应生成相应Cl-PAHs, 且温度升高有利于Cl-PAHs的生成.Yasuhara等[33]研究发现, 垃圾中所含的氯化钠和氯化铜等无机盐也可在氯化过程中提供氯源.此外, 有研究表明, PAHs的光致氯化也是Cl-PAHs形成的主要机制.Nilsson等[34]在紫外光照射条件下, 以CCl4溶液为氯源, 证实了PAHs的光化学氯化反应生成Cl-PAHs, 这是因为紫外光照射条件下产生的Cl-PAHs比无光条件下多.在进一步的研究中还发现, Cl-PAHs在炭黑或气溶胶上的生成速率要高于溶液中的生成速率[35].

改自文献[33, 34, 36] 图 3 燃烧过程中氯代多环芳烃的生成机制 Fig. 3 Formation mechanism of chlorinated polycyclic aromatic hydrocarbons during combustion process

综上可知, SPAHs可通过汽车尾气排放、生物质燃烧以及工业生产活动等形式一次生成, 还可通过母体PAHs在光化学反应下与空气中自由基生成、以及在热反应作用下二次生成.目前, 广泛研究了SPAHs的来源和生成方式, 但对于其在不同环境下生成路径和机制以及环境因子对该类污染物生成途径是否有影响等这些方面仍不清楚, 需要深入研究, 所以有必要进一步探究其在不同渠道下的反应作用机制和迁移转化行为.

2 SPAHs在不同环境介质中的存在状况与分布特征

目前, 多个国家和地区对大气、水体和沉积物中OH-PAHs、nitro-PAHs和Cl-PAHs等SPAHs的环境存在状况与浓度水平进行了报道.对此, 本文进行了总结, 如表 1所示.与PAHs类似, SPAHs已成为全球存在的一大类持久性有机污染物.虽然极地SPAHs的存在状况还不十分清楚, 仅检索到北极大气中存在nitro-PAHs[37], 但由于PAHs在极地广泛存在并能转化成SPAHs, 所以这些SPAHs潜在存在于极地环境中.

表 1 不同国家地区环境介质中SPAHs的浓度水平 Table 1 Concentration levels of SPAHs in environmental media from different countries and regions

2.1 大气中SPAHs

如表 1所示, 已有许多国家和地区开展了大气中SPAHs污染状况的研究, 这些SPAHs主要包括OH-PAHs、nitro-PAHs和Cl-PAHs.其中, 对于OH-PAHs, Simoneit等[42]分析了科瓦利斯(美国)、北京和太原冬季大气中13种OH-PAHs.北京和太原大气中ρ(总OH-PAHs)平均值分别为242 ng·m-3和417 ng·m-3, 在科瓦利斯未检出OH-PAHs.羟基芴为北京和太原两个城市大气中OH-PAHs的主要组分, ρ(总OH-PAHs)平均值分别为39.7 ng·m-3和72.3 ng·m-3.此外, OH-PAHs还在香港、南京和广州等大城市的大气中被检测到[42].

前人报道了北京、上海、西安和日本等地大气中的nitro-PAHs浓度水平, 如表 1所示.大气中的nitro-PAHs浓度水平主要与汽车燃料、地理位置和季节有关, 相对分子质量低的nitro-PAHs主要存在于气相中, 而相对分子质量高的nitro-PAHs主要存在于颗粒相中[70].Wada等[23]从日本长崎市大气颗粒物样品中发现, nitro-PAHs浓度与车流量成正比, 硝基化产生nitro-PAHs的反应物NOx主要来自于柴油车的排放, 柴油车产生的NOx是汽油车的5倍.Murahashi等[71]在主要由柴油机排放影响的大气颗粒物中检测出的硝基PAHs有1-硝基芘和二硝基芘(1, 3-二硝基芘、1, 6-二硝基芘和1, 8-二硝基芘).

除了车辆排放, 季节性变化也是影响nitro-PAHs浓度分布趋势的重要性因素.温度差异主要在冬夏两个季节, 冬季寒冷, 大气中nitro-PAHs的浓度会随着燃料使用量的增加而上升.另外, 由于冬季地面对流较为平缓, 受逆温现象的影响, nitro-PAHs会延长其在大气中的停留时间.而夏季因为阳光照射, 大气扩散较好, 污染物发生稀释或降解, 因此夏季相比其它季节, nitro-PAHs浓度水平要低.Marino等[72]从雅典春、夏、秋、冬这4个季节的大气颗粒物样品中检测到nitro-PAHs, 其浓度平均值范围分别为0.02~0.06、0.01~0.02、0.03~0.06和0.08~0.21 ng·m-3, 呈现出冬季>秋季>春季>夏季的分布趋势.

国内外学者对大气中Cl-PAHs的研究主要集中在东亚地区.Ma等[73]检测了上海郊区吸附在大气颗粒物PM2.5与PM10中的Cl-PAHs的含量, 研究发现Cl-PAHs普遍存在于大气可吸入颗粒物中.Jin等[74]发现煤燃烧的增加使得供暖期间北京大气中Cl/Br-PAHs的浓度平均值比非供暖期间高3~9倍, 这与Kakimoto等[51]研究的结果一致.Ohura[75]检测了日本城市名古屋一年四季中大气颗粒吸附态Cl-PAHs的浓度, ρ(总Cl-PAHs)年均值为43.3~92.6 pg·m-3; 在检测到的Cl-PAHs中, 1-ClPyr、7-ClBaA和6-ClBaP浓度比较高, 它们占Cl-PAHs总体的61%~76%.进一步, 还发现日本12个地区大气中的Cl-PAHs呈现出气相浓度在夏季较高, 而颗粒相浓度在冬季较高的季节变化规律[75].由此, 推测不同季节Cl-PAHs的来源可能存在差异, 其夏季来源可能是废弃物焚烧, 而冬季来源可能是车辆尾气排放和燃煤.此外, Kakimoto等[51]认为不同季节的光化学降解也会导致Cl/Br-PAHs浓度的季节性差异.另外, Helm等[76]在北极地区大气中检测出一定含量的多氯萘, 证实了Cl/Br-PAHs具有较强的大气远距离迁移传输能力.

2.2 水体和沉积物中SPAHs

地表水体中PAHs的主要来源有工业及城市废水的排放、地表径流、土壤淋溶、大气颗粒物的干湿沉降和水气交换等[77].而SPAHs和PAHs存在类似的来源, 所以SPAHs会以相同方式进入水体, 最终以溶解于水、吸附在颗粒物或呈乳化状态存在于水体中.水体中OH-PAHs来源广泛且成分复杂, 其浓度水平和分布特征研究较少.Wang等[78]检测了珠江口表层海水和底层海水中的1-萘酚、2-萘酚、2-羟基芴、3-羟基芴、9-羟基芴、2-羟基菲和3-羟基菲等12种OH-PAHs, 发现大部分OH-PAHs均能被检出, 总浓度范围为0.0387~1.0741 μg·L-1.目前OH-PAHs在水体中的监测研究仍处于起步阶段, 相关数据信息十分缺乏.

对于nitro-PAHs, 早在20世纪90年代, 马明生等[79]首次在我国城市饮用水和河水中检出了痕量的nitro-PAHs, 浓度范围为0.027~4.6 ng·L-1, 其中包括2-硝基萘在内的8种nitro-PAHs.虽然在水体中检出了nitro-PAHs, 但这并没有得到重视.直到2001年, Murahashi等[80]在日本金泽市Asano river和海水中均检出了ng·L-1水平的1-硝基芘, 才引起广泛关注.最近, Tiwari等[58]通过对我国太湖流域湖水、工业废水和市政污水中nitro-PAHs赋存水平的研究, 发现湖水中nitro-PAHs的浓度范围(<LOD~0.133 ng·mL-1)远低于工业废水(<LOD~1.004 ng·mL-1)和市政污水(<LOD~0.930 ng·mL-1)中的nitro-PAHs, 表明工业废水和市政污水可能是太湖湖水中nitro-PAHs的来源之一.

对于Cl-PAHs, 其分布主要取决于污染源种类及迁移转化过程(生物降解、光降解和吸附解析等), 某些Cl-PAHs同系物如6-ClBaP和1-ClPyr可作为其人为活动源的指示物[81].Shiraishi等[82]在日本筑波地区自来水样本中检测出萘、菲、芴和荧蒽的一氯/二氯衍生物, 但在地表水中未检测出, 推测自来水中的Cl-PAHs主要来源于氯化消毒及管道传输过程.然而, Wang等[60]在平湖中检测到大多数Cl-PAHs, Cl-PAH浓度范围在6.9 (9-ClPhe)~25.7 ng·L-1(1-ClPyr)之间, 并且高于自来水中的相应Cl-PAHs浓度含量.

水体中沉积物是SPAHs类化合物的重要归趋.余刚等[83]研究表明, nitro-PAHs分子结构仍然保留共轭体系, 其lgKow较母体PAHs (lgKow值为3.37~6.75)有一定的降低, 但仍具较强的疏水性, 如1-硝基萘、9-硝基蒽和1-硝基芴的lgKow分别为2.98、4.21和3.99.因此, 沉积物是nitro-PAHs在水体中的重要分布相.Witter等[84]对美国宾西法尼亚州中南部城市河流沉积物中nitro-PAHs进行了检测, 在35份样品中, nitro-PAHs检出率为100%, 含量为2.1~10.5 ng·g-1.Sankoda等[85]在日本Ariake Bay潮滩沉积物中检测到Cl-PAHs, 其含量在7×102~6.1×103 pg·g-1之间, 其中2-ClAnt (0.4~1.6 ng·g-1)和9, 10-Cl2Ant (0.27~1.33 ng·g-1)的含量最丰富.

2.3 土壤中SPAHs

相对于OH-PAHs, 国内外对于土壤中nitro-PAHs和Cl-PAHs的研究较多.对于nitro-PAHs, 有研究在西班牙城市土壤中监测到了4种nitro-PAHs[61], 含量范围为1.9~33 ng·g-1, 而De Guidi等[86]在意大利的土壤中检测到了小于1 ng·g-1水平的nitro-PAHs.在国内, Li等[87]研究发现, 西安表层土壤11种nitro-PAHs的含量为(118±52) ng·g-1, 主要成分为6-硝基苯[a]芘、2, 7-二硝基芴和5-硝基荧蒽, 其含量依次降低.Cai等[62]在长江三角洲地区11个城市的表层土壤中检测到了4种nitro-PAHs, 含量范围为0.4~4.6 ng·g-1.对于Cl-PAHs, 前人主要关注了其在污染源周边土壤中的存在状况.Tue等[88]发现电子废物露天焚烧可造成土壤的Cl/Br-PAHs污染, 且Cl-PAHs浓度比相应Br-PAHs高2.4~64倍.而化学工业园区(包括燃煤型热电厂、焦化厂、PVC制造厂和氯碱厂, 88 ng·g-1)和废弃电子垃圾回收厂土壤(26.8 ng·g-1)中的Cl-PAHs总含量平均值分别是农田表层土壤(0.15 ng·g-1)的590倍和180倍[50].

2.4 生物体中SPAHs

生物体中OH-PAHs主要作为PAHs的暴露生物标记物, 进行研究.Kammann等[68]分析了冰岛、法国、北海、德国湾和Baltic西部的不同种类鱼的胆汁中的1-羟基芘、1-羟基菲和3-羟基苯并芘, 其浓度范围为<LOD~654 ng·mL-1.钟林仁[69]采用GC-MS法分析了渤海湾大沽口和北塘口等地的海洋生物样品中的9-羟基菲.结果表明, 9-羟基菲在所有生物样品中均能检出, 且在生物体内各个器官中分布不均, 内脏中9-羟基菲浓度远高于别的器官, 北塘口梭鱼肠中的含量最高, 达283 ng·g-1, 最低的为北塘口梭鱼头, 只有0.04 ng·g-1.肌肉与头部OH-PAHs的浓度水平较低, 表明PAHs经代谢, 不容易在肌肉与头中富集.

Uno等[89]首次研究了海洋生物体内的nitro-PAHs, 测得在日本大阪湾贻贝体内的浓度ω(nitro-PAHs)范围为2.38~24.688 ng·g-1.Huang等[65]测定了密歇根湖中鲑鱼体内的nitro-PAHs, 共检出9种nitro-PAHs, 总浓度范围为0.2~31 pg·g-1.与水相中相类似, 1-硝基芘和6-硝基也是主要的污染物, 说明nitro-PAHs能在水相-生物体间进行传输[65].

Hattori等[17]分析了New Bedford港的蓝贝脂肪中的Cl-PAHs, 含量范围在14~28 ng·g-1之间, 并且在Cl-PAHs单体中, 1-ClPyr含量最多.有研究结果还发现, 海底底泥中Cl-PAHs的浓度远高于贻贝类生物中的浓度, 生物体内主要检测到相对分子质量低的Cl-PAHs, 而相对分子质量高的Cl-PAHs主要分布在底泥中, 生物利用性较差[17].Tan等[90]在辽宁省的22个淡水鱼样品中检测出3种相对分子质量低的Cl-PAHs (9-ClPhe、1-ClPyr和5-ClAce), 其中工业区附近的鱼样中Cl-PAHs的浓度较高, 所有的鱼样体内均以低分子量化合物为主, 这可能与其生物选择性有关.在已有研究中[90], 并未观察到Cl-PAHs与母体PAHs的相关性.

根据以上研究, SPAHs广泛存在于世界多个国家和地区的大气、水体和土壤等环境多介质中.对于SPAHs总的分布情况, Yan等[14]提出了总含量为：交通站点>城区>郊区>农村地区>偏远地区, 这也表明了SPAHs的主要排放源来自交通站点和城区.

3 SPAHs的光化学转化行为 3.1 SPAHs光降解动力学与反应路径

有研究表明, 光化学降解是PAHs及其衍生物SPAHs在环境中的重要消减方式[91].揭示PAHs和SPAHs的环境光化学行为, 对于该类污染物的环境归趋和生态风险评价具有重要意义.PAHs的光化学转化路径有羟基取代、光致氧化、光致氯化和硝基取代等, 生成毒性更强的产物, 表现为光致毒性效应[5].而SPAHs具有PAHs的母体结构, 其光化学行为也表现出多路径和高风险的特点[7, 92].

3.1.1 光降解反应动力学

从分子结构上分析(图 1), 助色团—OH、—Cl和生色团—NO2的引入, 使其与PAHs母体相比, 紫外-可见吸收光谱发生红移, 且相同波长下的摩尔吸光系数增加, 所以SPAHs对日光的吸收程度更强[5].OH-PAHs与nitro-PAHs等典型SPAHs的UV-Vis吸收光谱如图 4所示, 其在290 nm以上均有明显的光吸收.对于OH-PAHs, 2-羟基芴(2-OHFL)、9-羟基芴(9-OHFL)、1-羟基芘和9-羟基菲在模拟太阳光照射(λ>290 nm)下, 均可以吸收光子, 发生直接光解, 遵循一级反应动力学[5].相应的方程式如下：

(1)
图 4 典型OH-PAHs与Cl-PAHs的紫外-可见吸收光谱 Fig. 4 UV-Vis absorption spectra of typical OH-PAHs and Cl-PAHs

式中, c为污染物在t时刻的浓度; c0为初始(t=0)浓度; k为光解反应速率常数.继而, 采用对硝基苯甲醚/吡啶(PNA/Py)或2-硝基苯甲醛作为化学露光计(a), 可以测定底物s (如OH-PAHs)的量子产率(Φs), 计算公式如下：

(2)

式中, Lλ为光源在波长λ处的光强; ελ为化合物或露光计在波长λ处的摩尔吸光系数; Φa为露光计的量子产率.表 2总结了SPAHs光化学反应的动力学参数和量子产率(Φs).对比这3类SPAHs可以看出, Cl-PAHs的Φs(数量级为10-6)明显小于OH-PAHs的Φs (10-3~10-1)和nitro-PAHs的Φs (10-2).针对OH-PAHs, Ge等[5]研究了冰中4种模型化合物的光化学降解, 发现9-羟基菲和1-羟基芘的光降解量子产率(Φ)比冰雪中相应PAHs的Φ大一个数量级.并且, 将数据外推, 借助室外验证实验, 得到南北极夏天中午冰雪表面4种OH-PAHs的光降解t1/2为0.08~54 h.继而, 通过对比相同光照条件下冰中与水中OH-PAHs的光降解动力学, 发现冰中2-OHFL和9-OHFL的光降解分别比同等溶液水相中快3.0和7.5倍, 这归因于水溶液冷冻成冰过程中溶质(OHFLs)的冷冻浓缩效应.

表 2 不同条件下SPAHs的光化学反应速率常数(k)、半减期(t1/2)和量子产率(Φs) Table 2 Photochemical reaction rate constants (k), half-lives (t1/2), and quantum yields (Φs) of SPAHs under different conditions

对于nitro-PAHs, Xue等[93]考察了1-硝基芘和1, 8-二硝基芘的光解动力学, 发现汞灯照射下模拟降雪中, 这两种模型化合物均可发生快速消减, 反应遵循一级反应动力学, t1/2为1.55~1.90 h.此外, 对非水介质中或表面nitro-PAHs的光降解也有研究(表 2), 例如有研究者分析了1-硝基芘在玻璃片上的光解, 证实其光解符合二级动力学方程, 光解产物是羟基芘、二羟基芘和芘二酮等[94].

对于Cl-PAHs, 其在有机溶剂[95]和固体表面[96]的光化学行为有一些报道.为了模拟气溶胶表面Cl-PAHs的归趋, Ohura等[95]研究发现环己烷中11种Cl-PAHs光解速率常数(k)变化较大, 其中氯菲和6-氯芘的k比相应PAHs大.Niu等[96]运用QSPR模型预测了大气颗粒物上17种Cl-PAHs光降解半减期(t1/2, 0.56~20 h), 并指出第二最高占据轨道能等4个结构参数是影响t1/2的最主要因素.以上研究较好地阐明了非水环境中Cl-PAHs的光化学行为.

3.1.2 光转化路径与风险

对于OH-PAHs, 其在冰相或水相中光降解产物与路径的研究较为透彻.Ge等[5]鉴定了冰中4种OH-PAHs的转化产物, 并参考PAHs的光解机制[100, 101], 判断冰中OH-PAHs的光反应类型包含直接解离或被活性氧物种(ROS)氧化, 所涉及的转化路径为脱氢氧化、异构化和苯环羟基化.水相中OH-PAHs的光降解研究进一步发现, OH-PAHs比PAHs可能更容易被ROS氧化[5, 102], 例如9-OHFL可发生·OH参与的自敏化光解, 对表观光解反应的贡献是14.5%.Ge等[97]进一步采用衍生化技术和GC-MS/MS分析, 比较研究了冰中和水中9-OHFL的光降解产物和转化路径, 发现水中易于发生多羟基化(multi-hydroxylation)反应, 生成二羟基或三羟基代的产物, 而在冰中只是检测到单羟基化的产物(图 5).

改自文献[13, 97]; 转化产物标记为“TPn”, n表示相对分子质量 图 5 冰相和水相中9-羟基芴光降解产物和路径 Fig. 5 Photodegradation products and pathways of 9-OHFL in glacial and aqueous phases

相对于OH-PAHs, nitro-PAHs和Cl-PAHs的光降解产物与路径研究仅有少数报道.例如, Xue等[93]使用UPLC-TOF和GC-MS检测雪中1-硝基芘和1, 8-二硝基芘的光解产物, 但并未检测到转化产物, 推测这可能是因为中间产物在离子源中不易被离子化, 或者被气化.Kang等[103]发现水中1-氯萘和2, 3-二氯萘发生了脱氯、羟基化等光化学反应.除了表观光解, 在腐殖酸、NO3-/NO2-和Fe(Ⅲ)等光化学活性物质存在时, PAHs和SPAHs还可能发生间接光解[93], 但SPAHs的间接光解产物和路径未见报道.

污染物光化学转化过程中, 可通过光修饰或光敏化作用而产生光致毒性, 即光转化为毒性更大的产物, 或光敏化生成ROS, 从而对生物体造成损伤.有研究发现, SPAHs具有光致毒性.例如, 水中和冰中OH-PAHs光转化生成了对发光菌(Vibrio fischeri)毒性更大的产物, 表现为光修饰毒性[5, 9]; 冰中2-和4-氯酚对Vibrio fischeri也表现为光修饰毒性[104].水中nitro-PAHs受光照时生成ROS, 对日本虎斑猛水蚤(Tigriopus japonicus)生物体造成损伤, 表现为光敏化毒性[7]; 蒽醌类化合物对大型溞(Daphnia magna)也具有光敏化毒性[105].以上研究表明, SPAHs类化合物的光致毒性效应值得关注, 有待深入研究.

3.2 SPAHs光降解的影响因素及作用机制 3.2.1 不同溶剂对SPAHs光降解的影响

大气中的SPAHs往往由母体PAHs与·OH、氮氧化物和氯离子反应形成, 随后分配到颗粒相以气溶胶的形式存在, 再以干湿沉降和大气扩散等方式进入水体和土壤等介质中.推测介质通过多种作用方式影响SPAHs的光解, 有研究者选取有机溶剂来反映介质的内在性质, 考察其对SPAHs光解的影响机制.有研究通过比较1-硝基芘在多种有机溶剂中光降解和产物形成的量子产率[106], 发现其量子产率与溶剂类型有显著的依赖性(表 3).1-硝基芘在苯和异丙醇的产率较高, 另外并有产物是由苯氧通过自由基重排形成.另外, 在模拟气溶胶介质中, 酚类物质可加速1-硝基芘的光消减, 并且能够得到较大的产物生成量子产率[106].

表 3 不同溶剂中1-硝基芘光降解和产物形成的量子产率 Table 3 Quantum yields of photodegradation and product formation of 1-nitropyrene in different solvents

3.2.2 水环境因子对SPAHs的光化学行为的影响

水体性质和溶解性物质会影响污染物的化学光降解, 包括pH、溶解性有机质(DOM, 包括腐殖酸HA等)、NO3-/NO2-和Fe(Ⅲ)等.其中, HA是天然水体中影响光降解的重要因素, 其光化学过程及其引发的有机污染物(P)的转化如图 6所示.在不同的光源照射下, HA可对污染物的光降解表现为促进与抑制两种作用[107].

图 6 腐殖酸(HA)的光化学过程及其引发的有机污染物(P)的转化 Fig. 6 Major photochemical processes of humic acids (HA) that lead to the transformation of organic pollutants (P)

对于SPAHs, 冰雪环境因子对其光降解影响的相关报道非常少, 仅检索到单一溶解性物质对光解动力学的影响.H2O2和NO3-的存在加快了雪中硝基芘的光降解[93]; 硝酸盐和铁离子加快了1-氯萘和2, 3-二氯萘的光化学转化反应[103].Cl-、NO3-、Fe(Ⅲ)和低浓度腐殖酸均可促进9-OHFl的光降解, 这可以初步解释海水冰、淡水冰中光降解较纯水冰中快.Ge等[5]研究了水环境中主要溶解性物质对OH-PAHs光降解动力学的复合效应, 发现自然水体中9-OHFl的光解速率比纯水中慢.这表明不同冰和不同水中9-OHFl光解动力学的规律正好相反, 推测其归因于冰和水中主要溶解性物质光化学活性的差异.进一步, 通过比较研究冰中和水中主要溶解性物质对OH-PAHs 2-羟基芴(2-OHFL)和9-羟基芴(9-OHFL)光化学转化动力学的影响, 发现不同相中某一溶解性物质对光降解动力学的影响作用并不完全一样, 冰相和水相Cl-表现为相反的效应; 两相中腐殖酸钠盐(HASS)抑制光降解, NO3-和Fe(Ⅲ)对光降解的作用存在一些差异(图 7)[97].以上结果表明, 相同光照条件下, 这些溶解性物质的水环境光化学和冰雪环境光化学存在差别[97].

改自文献[97] 图 7 模拟日光照射下, 冰/水中溶解性物质对2-羟基芴(2-OHFL)光降解速率常数(k)的影响 Fig. 7 Effects of aqueous constituents on photodegradation rate constants (k) of 2-hydroxyfluorene (2-OHFL) under simulated sunlight irradiation

4 展望

(1) 多介质环境中SPAHs的分布特征和分配规律需要深入研究.目前, SPAHs在单一介质中的存在状况多有报道, 但在复杂环境基质中的赋存状况、时空变化规律需要深入研究; 并且, 大气-水体-沉积物-生物等多介质间的SPAHs分配规律还鲜有报道, 需要探索.以上研究工作, 可为新污染物治理行动中“筛”和“评”提供必要的基础数据.

(2) 大气和水/冰中SPAHs的光化学行为机制值得探索.环境多介质中SPAHs分布广泛, 水或冰雪中SPAHs的模拟光化学行为有一些报道.然而, 以SPAHs为模型化合物, 关键的光化学转化机制, 如环境因子复合效应、自由基反应机制、同位素分馏, 还未见报道.深入探索SPAHs的转化过程和机制, 也是今后控制和消减该类新污染物面临的迫切问题.

(3) 环境中SPAHs的风险和危害需要准确评估.环境中SPAHs存在着很大的“储存库”, 某些SPAHs毒性高于母体, 或表现为光致毒性.该类新污染物具有来源广泛性、危害严重性、风险隐蔽性、环境持久性的多重特点, 其风险值得重点关注.因此, 有必要采用标准受试生物暴露实验或计算毒理学筛查技术, 获取更多SPAHs系列物的毒性参数, 为其健康风险评价提供支持, 并为该类新污染物的源头治理提供科学依据.

5 结论

多环芳烃衍生物(SPAHs)是一类来源广泛且具有较大环境风险的新型有机污染物.本文通过总结OH-PAHs、nitro-PAHs和Cl-PAHs的环境存在状况与光化学行为的最新进展, 讨论了这3类SPAHs的环境来源、形成机制和分布特征, 并重点评述了水和冰等介质中SPAHs光化学转化动力学、反应路径和影响因素.相对于传统污染物PAHs, 新污染物SPAHs的研究深度和广度还远远不够, 许多科学问题需要探索或研究.

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