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多环芳烃（polycyclic aromatic hydrocarbons，PAHs）是一类由2个或2个以上苯环以线状、角状或簇状排列组合而成的稠环化合物，具有致畸性、致癌性和致突变性. 它们是一类典型的持久性有机污染物，广泛分布于不同的环境介质中. PAHs的来源有自然来源和人为来源，在许多受人类活动影响的地区，PAHs的污染主要是人类活动的结果，包括化工燃料的不完全燃烧，石油化工厂、炼油厂、焦化厂等工业污染排放，垃圾焚烧，海上交通运输以及石油泄漏等.
PAHs的种类很多，目前对于PAHs的研究不再局限于母体多环芳烃（parent polycyclic aromatic hydrocarbons，P-PAHs），同时还包括各类毒性更强、更难降解的衍生物，包括烷基化多环芳烃（alkyl PAHs，A-PAHs）、卤代多环芳烃（halogenated PAHs，H-PAHs）、硝基多环芳烃（nitrated PAHs，N-PAHs）和杂环多环芳烃（heterocyclic PAHs）. 一些PAHs的衍生物已被证实具有比P-PAHs更强的毒性且更难以被生物降解[1]. 因此研究PAHs及其衍生物的污染特征，对这一大类持久性有机污染物的风险管控及污染修复，具有重要的环境意义. 海洋被认为是各种来源有机污染物最终的汇[2]. 在人类活动和气候变化多重压力下，海洋中持久性污染物的生态风险呈持续增加趋势，长期积累有可能加快海洋生态系统的退化. 本文对海洋中P-PAHs及各类衍生物的污染特征和来源进行归纳总结，为海洋中PAHs的风险评价提供更全面的数据支撑.
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P-PAHs是指没有任何取代基和杂环原子的芳烃化合物，在自然环境中普遍存在，主要由不完全燃烧产生，如固体燃料燃烧和车辆排放等. 与其衍生化合物相比，P-PAHs在海洋环境中的浓度更高[3]. 据报道，许多致癌性、免疫和生殖疾病都与暴露于P-PAHs有关[4]. 美国环保局（USEPA）在上世纪80年代就把16种P-PAHs确定为环境中的优先控制污染物，我国也把16种P-PAHs列入环境优先监测的污染物黑名单.
P-PAHs广泛分布于海洋的水体环境中. 海南省东寨港红树林水体旱季检测到的14种P-PAHs总质量浓度（∑14P-PAHs）范围为1015.27—2069.07 ng∙L−1，雨季的浓度为234.46—683.33 ng∙L−1 [5]. 广西廉州湾和三娘湾夏季16种P-PAHs的总质量浓度（∑16P-PAHs）在河流的浓度为57.89—90.76 ng∙L−1，高于海岸带水体的浓度（49.75—68.65 ng∙L−1）；冬季入海河流和海岸带水体的∑16P-PAHs分别为（106.67±29.96）ng∙L−1和（92.43±22.19）ng∙L-1[6]. 由此可见，海洋水体中P-PAHs的浓度均呈现出旱季大于雨季、冬季大于夏季的规律，可能是因为雨季降雨量大，对水体中的PAHs起了稀释作用，强紫外线促进PAHs光解的原因[7]. 另外，夏季高温的环境也利于一些PAHs降解菌的繁衍和生长，使降解和吸附于菌内的PAHs含量增加[8].
国内外不同海域表层沉积物中P-PAHs的总浓度汇总于表1. 从全国范围看，珠江和长江流域作为我国典型流域，P-PAHs浓度处于较高的污染水平. 珠江口流域的P-PAHs污染水平高于长江流域，这与珠江流域沿岸更发达的工业相关. 东海在国内外海岸的P-PAHs污染中整体处于轻度污染水平. 根据质量基准法，三亚河表层沉积物对生态环境并未产生负面影响，处于低风险水平[10]. 细颗粒的黏土和粉砂是内陆架沉积物的主要组成部分，沉积在河口处和浙闽沿岸的沉积物分别占长江入海泥沙的40%和32%[17]. 东海东面的陆架表层沉积物中和长江河口处的P-PAHs的含量最高，说明陆源汇入对陆架表层沉积物中P-PAHs的积累起主要作用[12]. 对国外海域，韩国沿海表层沉积物中P-PAHs的浓度最高，远高于东北大西洋浅滩及西北大西洋伊比利亚半岛. 由于韩国沿海的经济活动强度远大于东北大西洋浅滩及西北大西洋伊比利亚半岛，说明海洋环境中P-PAHs的浓度水平与人类活动密切相关.
根据同分异构比值法，沉积物中P-PAHs主要由3—5环组成，占总P-PAHs的84.4%[18]. 大部分海域检测到的P-PAHs都以3环为主，占总量的三分之一以上；其次是4环及5环，6环最低[19]. 石油产品是沉积物中3环P-PAHs的主要来源，5环及6环P-PAHs主要来自化石燃料的不完全燃烧[20]. 陆源输入对沉积物中P-PAHs浓度的影响起主要作用，对于黄海以南中国东方的海域及其大陆架，其沉积物中的P-PAHs主要以3环为主，来源主要以航船泄露的石油为主[20]. 以三亚河为代表的穿越市区的河流，其沉积物主要以3—5环为主，来源以市区木柴和煤的燃烧为主；而对于黄河口、黄海等位于中国北方的流域的沉积物中的P-PAHs主要以6环为主，这可能是因为中国北方冬天气温太低，需要大量燃煤供暖的缘故[21].
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P-PAHs上的氢原子被烷基取代后形成A-PAHs，比相应的P-PAHs在热力学上更不稳定，在高温和岩源中分别占总多环芳烃（T-PAHs，包括P-PAHs与PAHs衍生物）的60%和99%[22-23]. 由于A-PAHs在不同环境中具有较高的亲脂性和丰度，其净毒性的比例高于相应的P-PAHs[24]. A-PAHs已被证明在鱼类胚胎中产生类似二噁英的毒性，如重烯（7-异丙基-1-甲基苯），烷基可以改变其毒性，并为羟基化提供额外的位点[1]. 在致癌性方面，蒽和苯并[a]蒽的烷基化形式比它们的非烷基化同系物更具致癌性[25]. 此外，A-PAHs可以转化为其他毒性更强的化合物，如N-PAHs[26]. 因此A-PAHs会对环境生态安全构成更大的威胁[27].
A-PAHs在海洋环境中的丰度很高. 本研究对中国南黄海、中国东海和国外部分海域表层沉积物中的A-PAHs的污染水平进行汇总，结果见表2.
从表2可以看出，不同海域的表层沉积物中A-PAHs的污染水平有很大的差异. 总的来说，中国南黄海的A-PAHs污染水平较高，巴西巴拉那瓜河口的污染水平最低，呈现出近海的污染水平高于外海的趋势. 另外，本课题组对珠江口、大亚湾和南海北部的表层沉积物中60种PAHs（包括31种P-PAHs和29种A-PAHs）进行了全面的调查分析，发现A-PAHs在珠江口、大亚湾和南海北部的浓度范围分别为115—766、127—354、200—272 ng·g−1，占60种PAHs的32%—36%，在各A-PAHs中，浓度最高的为烷基萘和烷基菲[33]. 从浓度水平来看，我国珠江口的A-PAHs污染水平远高于中国南黄海、东海和国外部分海域，大约高2—3个数量级.
珠江口不同采样点之间A-PAHs浓度水平差异大，原因可能是悬浮颗粒物通过淡水随河流输入，大部分沉积在三角洲的北部，导致PAHs的浓度由北向南、由东向西呈下降的趋势，该分布趋势与珠江口、大亚湾和南海的总有机碳（TOC）的含量分布一致[33]. 其他研究者也发现，沉积物中有机污染物浓度与TOC含量有显著的正相关关系[34]. 然而对于沉积物中的复杂有机质，TOC的哪一种组分在控制自然沉积物中PAHs的分布起关键作用尚不清楚，还需要进一步研究. 此外，南海表层沉积物中A-PAHs对于T-PAHs的占比小于珠江口，原因可能是南海受人类活动和河流输入的影响小，因此珠江口沉积物中的A-PAHs浓度高于大亚湾和南海北部，说明人类活动和河流径流对于A-PAHs的产生和转化起重要作用[33]. 因此有必要量化河流径流影响下，有机污染物从沿海地区向海洋的运输和分布模式，这对于了解有机污染物在全球范围内的污染特征具有重要价值.
从石油污染地区采集的生物样品，在鳍鱼和贝类肌肉组织中检出A-PAHs[35]. 尼日利亚东南部虾样品中A-PAHs的浓度为(31.38±18.49) ng·g−1 ww（ww: wet weight, 湿重），对应沉积物中A-PAHs的浓度为172.36 ng·g−1 dw（dw: dry weight，干重）[30]. 说明海洋生物体内的A-PAHs与该海域表层沉积物中的A-PAHs浓度有很大的相关性. A-PAHs的地理分布和径流模式有明显的相关性，可推测近海石油活动是当地海洋生物体内的A-PAHs的重要来源. 位于发达工业化国家日本的东京湾被检测到无论是水体、沉积物还是海洋生物都含有更高浓度的A-PAHs，这点与国内珠江口PAHs的污染情况相似[36].
与非烷基化的PAHs相比，A-PAHs的分布也可以提供关于这些污染物的来源信息. A-PAHs在热力学上不像对应的P-PAHs那样稳定，因此，燃烧通常会导致非烷基化PAHs相对于烷基化同系物的富集（例如，萘的浓度高于C2-萘，C2表示2个C上有烷基取代），但具体的模式随着燃烧温度的变化而变化[37]. A-PAHs的相对丰度可用来区分其来源，随着温度的升高，A-PAHs中烷基侧链的丰度降低；与木材燃烧产物相比，煤炭燃烧产物中A-PAHs的浓度相对较高；A-PAHs在低温（100—150 ℃）下形成，而在高温下（2000 ℃）形成的含量很低，P-PAHs在热力学上比A-PAHs更稳定[38]. 因此，在高温下形成的燃烧产物中，A-PAHs被耗尽，而在地壳中低温下成岩生成的石油中，A-PAHs含量丰富. 在原油污染环境中A-PAHs的浓度远高于P-PAHs. A-PAHs的空间分布具有地点特异性，表明存在石油污染源[39]. 通常，石油源PAHs以A-PAHs和较轻的PAHs为主，较重的PAHs较少，而热源多为高环PAHs[40]. 在雅加达湾，尽管PAHs剖面有丰富的烷基化和较轻的PAHs，但也有相当数量的较重的PAHs，这表明大部分工业化国家一样，热生成源的贡献更大，但也有岩石源的输入.
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H-PAH是由氯或溴原子取代PAHs碳骨架上的氢原子而形成的一类新污染物，如氯化多环芳烃（Cl-PAHs）和溴化多环芳烃（Br-PAHs）[41-42]. 它们在各种环境介质中广泛存在，在近年来引起了越来越多的关注[43]. 尽管它们的生产机制被认为是相似的，但是它们在环境中的含量可能比相应的P-PAHs低约10—100倍[44]. 与P-PAHs类似，H-PAHs具有致畸、致诱变和致癌性，因为它们可以结合并激活芳香烃受体（aryl hydrocarbon receptor，AhR）并诱导DNA损伤[45]. 有研究发现，Cl-PAHs与P-PAHs相比具有显著的DNA损伤作用[46]. Huang等[45]发现, 部分低芳环数的Cl-PAHs能引起比P-PAHs更强烈的AhR效应，具有类二噁英的毒性. H-PAHs的毒性与其空间尺寸相关[47]，低环Cl-PAHs诱导AhR的活性随着氯原子数量的增加而升高[23]. 除了引起DNA损伤作用，H-PAHs还具有免疫毒性. Li等[48]发现, Cl-PAHs在不激活AhR的较低浓度下（1 μmol·L−1）能诱导免疫毒性.
目前对国内外海洋水中H-PAHs的研究较少，可能与水中H-PAHs检出率较低有关. 我们团队对珠江口的H-PAH进行调查分析，发现水中H-PAHs的检出率很低，仅有1-溴芘被检出[49]. 目前对海洋悬浮颗粒物和沉积物样品中H-PAHs报道较多，尤其是表层沉积物. 国内外不同海域表层沉积物中H-PAHs的污染水平见表3.
本团队对珠江口的H-PAH进行调查分析，发现在表层沉积物中Cl-PAHs的浓度（600—25700 pg·g−1）高于Br-PAHs的浓度（800—66300 pg·g−1），其中9-氯菲为浓度最高的H-PAHs[49]. 珠江口的H-PAHs浓度远高于广西茅尾海和黄海. 从全球范围看，我国珠江口H-PAHs的浓度处于中等偏高水平. 东京湾沉积物岩心的Cl-PAHs和16个优先控制PAHs的通量分别为29—570 pg·（cm2·a）−1和85000—609000 pg·（cm2·a）−1，通量在10世纪50年代最低，在1989—1990年最高；Cl-PAHs遍布东京湾沉积物的整个沉积岩芯，浓度范围为36—1210 pg·g−1，其中最高浓度出现在沉积岩芯的14—16 cm节段，与20世纪90年代中期相当，这个剖面特征与以前报道的飞灰和城市空气样本相似；而对于个体Cl-PAHs，6-氯苯并[a]芘和1-氯芘是沉积物中的主要化合物，来源和分布与P-PAHs的来源和分布直接相关[52]. 新贝德福德港检测出的Cl-PAHs的平均浓度较低，为1880 pg·g−1 dw，而新贝德福德港贻贝中Cl-PAHs的平均浓度为21000 pg·g−1，比沉积物中的浓度高1个数量级[52].
在一项早期研究中，在自来水样品中检测到了10—100 pg·L−1水平的Cl-PAHs，Cl-PAHs的浓度从高到低排序为：9-氯菲（330 pg·L−1）> 9,10-二氯菲（180 pg·L−1）>3-氯芴（150 pg·L−1）> 2-氯芴（130 pg·L−1）；在湖水（水处理厂的原水）中，P-PAHs的浓度为380 pg·L−1，未检测到Cl-PAHs，说明自来水中出现的Cl-PAHs是通过与传统氯消毒产生的余氯反应在自来水供应系统中形成[53]. 海洋中存在的H-PAHs是否来自于自来水供应系统，这个问题值得进行溯源研究.
珠江口水体悬浮颗粒物上的H-PAHs的浓度范围为66000—1423000 pg·g−1，在悬浮颗粒物上的Cl-PAHs和Br-PAHs的浓度均与P-PAHs的浓度显著相关[49]，表明H-PAHs和P-PAHs通过悬浮颗粒物的集体转运. 然而，在沉积物样品中没有发现H-PAHs与P-PAHs之间的显著相关性[42]. TOC对P-PAHs等疏水有机而化合物的迁移和分布起着重要作用[54]. 珠江口H-PAHs浓度与TOC水平显著相关，这说明TOC水平对H-PAHs的贡献较大[49]. 根据已发表的文章，电子垃圾回收、汽车排放和垃圾焚烧被认为是环境H-PAHs的主要排放源[42]. 东京湾、新贝德福德港沉积物中H-PAHs的主要来源是水域附近的电子垃圾回收和垃圾焚烧[55]. 从不同来源释放的H-PAHs通常具有特定的成分，例如，含有两种或两种以上的Cl-PAHs通常来源于垃圾焚烧，在交通尾气中检测到含有少量的Cl-PAHs[56]. 故可根据具有特征的H-PAHs进行溯源和源头治理.
目前，H-PAHs产生机制的细节尚不清楚. 此外，通过食物链进行的生物积累和/或生物放大的潜力及其进行长期大气运输的能力还需要更多的研究来进行评估. 同样，人类通过各种途径接触H-PAHs的细节仍不清楚. 此外，高取代的H-PAHs（P-PAHs中超过3个卤原子）和ClBr-PAHs（同一P-PAHs中的氯和溴原子）的环境发生和行为值得关注，其产生机制、环境迁移转化规律、人类暴露和健康风险是未来研究的重点.
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N-PAHs是由硝基取代PAHs碳骨架上的氢原子的PAHs衍生物，它在PAHs的芳香族苯环上至少有一个硝基官能团[57]. N-PAHs以其致癌性和致突变性闻名[58]. 作为PAHs的硝基衍生物，N-PAHs因其具有比PAHs更强的诱变性而成为人们关注的焦点[59]. 部分N-PAHs的致癌性可达其P-PAHs的10倍，致突变性则可达其P-PAHs的105倍[60]. N-PAHs具有诱变性、遗传毒性、致癌性和雌激素活性，经过富集后对水生生物和人类的威胁均不可忽视. 根据一项已发表的科学研究，人类致癌风险评估工作组已将1-硝基萘、2-硝基萘、7-硝基苯并[a]蒽等N-PAHs列为致癌组，其中一些N-PAHs已经被赋予了毒性等效因子，其中一些值远高于苯并[a]芘[61].
大多数N-PAHs，如1-硝基芘（1-NP），主要是由化石燃料和生物质的不完全燃烧和热解产生[62]，而2-硝基芘和2-硝基荧蒽是通过大气反应形成的二级PAHs衍生物[63]. 大气中N-PAHs的一次和二次形成反应在不同的时间和地点差异很大. 由于它们的蒸汽压力不同，4环及以上的N-PAHs主要出现在粒子相中[64]. 与它们相应的P-PAHs相比，N-PAHs具有更高的分子量、辛醇-空气分配系数、粒气分配系数和更低的蒸汽压力、水溶性、辛醇-水分配系数、有机碳水分配系数和亨利常数[64]. 一般来说，P-PAHs上单个NO2-基团取代产生的N-PAHs，分别使蒸汽压力和水溶性降低3个数量级和1个数量级[65]. PAHs被硝基取代后，辛醇-水分配系数一般可降低25%—80%[66]. 它们与P-PAHs一起从车辆、工业、家用炉灶/加热器、垃圾焚烧炉和自然火灾中排放到环境[67].
由于N-PAHs的水溶性较低，故在水中检出率较低，报道较少，而N-PAHs在沉积物中的报道较多. 不同地区沉积物中N-PAHs总浓度水平见表4. 日本广岛湾与苏伊蒙河沉积物中N-PAHs总浓度较高，西班牙巴塞罗那海岸近海沉积物中N-PAHs浓度较低. 从整体浓度水平来看，海洋中N-PAHs的浓度高于A-PAHs和H-PAHs，但是比P-PAHs低2—3个数量级.
作为一类新污染物，N-PAHs在海洋中的污染调查还比较少，在废水及大气中的研究比较多. 在一项来自中国北京的废水处理的进水和出水的研究中，未检测到N-PAHs（2-硝基芴、3-硝基荧蒽、1-硝基芘、9-硝基苯和7-硝基苯并[a]蒽）[73]. 在日本加油站的36份废水样品（从油水分离罐中采集）中，1-硝基芘和1,6-二硝基芘的浓度分别为0.139 ng·L−1和0.002 ng·L−1，被认为是加油站废水中N-PAHs的主要来源[74]. 在连接日本中央高速公路的排水管中收集的街道径流中也检测到N-PAHs，溶解相中∑10N-PAHs的浓度为2.3—4.9 ng·L−1，而∑18P-PAHs浓度为34—160 ng·L−1，颗粒相的∑10N-PAHs和∑18P-PAHs浓度分别为11—73 ng·L−1和350—11000 ng·L−1，在街道径流中，81%—97%的N-PAHs被吸附在颗粒物中，N-PAHs占总PAHs的1%—30%[75]. N-PAHs多在大气颗粒中被发现，可能通过沉降进入水体，更深入的研究大多是针对大气中的N-PAHs，特别是在颗粒物中.
虽然N-PAHs在水体中能检测出的浓度较低，但是仍可对水生生物产生不可忽视的毒害作用. N-PAHs可被水生生物从受污染的水、悬浮颗粒、饲料或沉积物中吸收[76]. N-PAHs的lgKow为2—6时，有可能发生生物积累[65]. Steinberg等[78]通过一级动力学表明, 2-硝基芴可在大型溞（Daphnia magna）中生物积累. 密歇根湖收集的湖鳟鱼中全鱼和鱼卵中Ʃ9N-PAHs的平均浓度分别为7 pg·g−1和22 pg·g−1，检测到的致癌化合物包括1-硝基芘和6-硝基氯烯[77]. 在瑞典的海洋动物中发现9-硝基蒽在鱼类中的浓度高达2500 pg·g−1，在软体动物中高达2400 pg·g-1[79]. 多目鱼（Marbled flounder）饲料中N-PAHs的生物创伤因子（BMF）≤0.02，表明饲料中N-PAHs缺乏显著积累[80]. 另有研究发现，溶解在水中的N-PAHs会被生物累积，2-硝基芴、3-硝基菲、1-硝基芘和1,8-二硝基芘的生物富集因子（BCF）在4—422之间，BCF与lg Kow呈负相关关系，这与其他疏水有机化合物常见的结果相反，这些N-PAHs的BCF值明显低于与其对应的P-PAHs[81].
N-PAHs的形成和归趋不能仅靠人为活动的规模来预测，因为在这个阶段不能解释PAHs排放后转化对环境中总浓度的贡献. N-PAHs由于其土壤水溶性低、对土壤固相吸附高、渗入地下水的能力低，再排放为气相，固定在土壤中，所以目前对海洋环境中N-PAHs的分布与转化研究较少.
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杂环PAHs构成包含至少一个杂环的PAHs亚群，即具有除碳和氢以外的至少一个原子，这些化合物在环境中同时以被取代的形式和未被取代的形式出现，当PAHs的熔融环结构中的一个或多个碳原子被氮、硫或氧原子取代时，就形成了这些化合物[82]. 含氮、硫和氧杂原子的PAHs分别被称为氮杂芳烃（PANHs）、硫代杂芳烃（PASHs）和氧杂芳烃（PAOHs）. 杂环PAHs的来源与P-PAHs化合物相似，因此，它们经常与P-PAHs在环境中共存[83]. 杂环PAHs的主要人为污染源是化石燃料不完全燃烧过程中的大气沉积、石油泄漏和废水灌溉[84]. 氮、氧、硫原子的引入，导致PAHs的极性和水溶性显著增加，特别是对于PANHs来说，例如萘作为最简单的PAH，其水溶性约为30 mg·L−1，而相应的氮杂环多芳烃喹啉的溶解度为6100 mg·L-1[85].
Brinkmann等[86]确定了杂环PAHs的雌二醇等效因子：对于吖啶、氧杂蒽、茚、2-甲基苯并呋喃、2,3-二甲基苯并呋喃、二苯并呋喃、二苯并噻吩、喹啉和6-甲基喹啉，这些值与其他异种雌激素（如烷基酚或双酚A）相当，虽然受焦油污染的地点是杂环PAHs的已知来源，特别是焦化厂往往位于河流附近，但关于它们在海水水相甚至河流本身的数据很少. Shinohara等[87]在1983年测定了日本海水中的14种氮杂环PAHs，发现2,4-二甲基喹啉的浓度高达55 ng·L−1. 此外，一些杂环芳烃，如吖啶和吖啶酮被认为是抗癫痫药物卡马西平的转化产物[88]，西班牙埃布罗河及其支流中检测到这两种物质的浓度高达18 ng·L-1[89]. 杂环PAHs化合物普遍存在于焦化、石化、木材防腐剂和其他相关行业产生的废水中. 杂环PAHs是从类似于PAHs的来源被引入到环境中的，这使得很难确定它们的具体来源. 这导致这些化合物未被列入优先控制污染物清单，与P-PAHs相比，更高的极性提高了它们的生物有效性，而更高的毒性增加了其生态毒性和对人类的健康风险.
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存在于海洋中的众多种类的PAHs中，P-PAHs和A-PAHs的浓度占比最高，种类相对更丰富，对环境健康的威胁也更大. P-PAHs在热力学上比A-PAHs更稳定，近海石油活动是A-PAHs的主要来源. 在上述PAHs中，沉积物中所测得的种类和浓度均比水体中的PAHs更多更高，且河流输入对沉积物的影响较大，说明内陆城市的源头治理至关重要. 除了河流输入沉积物之外，石油也是海洋PAHs的重要来源之一. 石油污染中含有高比例的P-PAHs和A-PAHs. N-PAHs更多则是其对应的P-PAHs在大气中通过紫外线照射或与自由基反应生成，后随沉积物沉降到水体中[73].
PAHs及其衍生物通过食物链进行的生物积累或生物放大的潜力及其进行长期大气运输的能力还需要更多的研究来进行评估. 同样，人类通过各种途径接触PAHs的细节仍不清楚. 此外，高取代的H-PAHs、N-PAHs的环境行为值得关注. 因此，高取代PAHs的产生机制、环境迁移转化规律、人类暴露和健康风险是未来研究的重点.
若要对海洋环境中痕量的PAHs衍生物进行更广泛和更全面的监测，需要建立更灵敏、更高分辨率的分析方法. 高分辨率质谱分析将成为准确识别海洋环境样品复杂基质中痕量PAHs及其衍生物的有力工具. 建立快速、灵敏、特异、高通量的方法同时测定各种样品中的各类PAHs，以便准确评估PAHs及其衍生物对人体健康风险，是未来的研究趋势.
海洋多环芳烃及其衍生物的污染特征和来源分析
Characteristics and source analysis of polycyclic aromatic hydrocarbons and their derivatives in marine environment
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摘要: 随着经济的发展，人类生产活动产生的污染物不断增加，海洋成为所有污染物最终的“汇”. 多环芳烃因其种类繁多、毒性强、难降解、分布广泛而备受关注. 它们在海洋环境中被频繁检出，是海洋中常见的污染物. 本文对国内外海洋水体、沉积物及海洋生物体内母体多环芳烃、烷基多环芳烃、卤代多环芳烃、硝基多环芳烃和杂环多环芳烃的污染特征、来源及毒性进行归纳总结，并进行展望.Abstract: With the development of economy, the pollutants produced by human production activities continue to increase, and the ocean becomes the final "sink" of all pollutants. Polycyclic aromatic hydrocarbons (PAHs) have attracted much attention due to their variety, high toxicity, difficult to degradation and wide distribution. They are frequently detected in marine environment and are common pollutants in the ocean. In this review, the characteristics, sources and toxicity of parent PAHs, alkyl PAHs, halogenated PAHs, nitrated PAHs and heterocyclic PAHs in marine water, sediments and marine organisms at home and abroad were summarized, and the results were prospected.
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Key words:
- polycyclic aromatic hydrocarbons /
- persistent toxic pollutants /
- marine /
- sediments.
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表 1 海洋表层沉积物中P-PAHs的总浓度
Table 1. Total concentrations of P-PAHs in marine surface sediments
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表 2 海洋表层沉积物中A-PAHs的总浓度
Table 2. Total concentrations of A-PAHs in marine surface sediments
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表 3 海洋表层沉积物中H-PAHs的浓度水平
Table 3. Concentrations of H-PAHs in marine surface sediments
国家及地区
National and Location检测的H-PAHs数量
Number of H-PAHs analyzedH-PAHs总浓度/（pg·g−1）
Concentration of total H-PAHs参考文献
Reference中国珠江口 7 Cl-PAHs 600—25700 [49] 7 Br-PAHs 800—66300 中国广西茅尾海 18 Cl-PAHs 300—9600 [50] 中国黄海 20 Cl-PAHs 290—1130 [51] 11 Br-PAHs 6—248 斯里兰卡尼甘布 20 Cl-PAHs 320—1798 [51] 11 Br-PAHs 15—104 斯里兰卡康提 20 Cl-PAHs 552—2381 [51] 11 Br-PAHs 20—160 日本东京湾 20 Cl-PAHs 36—1210 [52] 美国萨吉诺河和萨吉诺湾 20 Cl-PAHs 49—2490 [52] 新贝德福德港 20 Cl-PAHs 1880 [52]
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