2024, Vol. 54
2. 山东省海洋资源与环境研究院，山东省海洋生态修复重点实验室，山东 烟台 264006
阿特拉津，作为一种三嗪类除草剂，因其低成本和出色的除草效果，已在全世界范围内广泛使用[1]。2018年，阿特拉津的全球销售额高达6.55亿美元，占三嗪类除草剂市场份额的46.8%[2]，在中国农业上的年使用量约为1 000~1 500 t[3]。尽管其效果显著，但阿特拉津在土壤中的半衰期较长，约为150~360 d，这意味着它有可能长期存在并污染我们的环境。更糟糕的是，这种化学物质在施用后会进入河流、湖泊等地表水，最终汇入海洋，成为海洋环境中的常见污染物[4]。在中国海州湾和澳大利亚大堡礁海水中，阿特拉津的浓度分别达到了61.9和730 ng/L[5-6]。另外，在中国海参样品中也检测到了阿特拉津，含量为0.9~3.62 μg/kg[7]。
随着阿特拉津在海洋环境与生物体内不断被检出，它对海洋生态系统的潜在危害引起了人们的持续关注，这种化学物质不仅对海洋生物产生直接毒性效应，而且通过食物链累积，对更高营养级的生物产生影响。例如：将三角褐指藻(Phaeodactylum tricornutum)暴露于3.2 μg/L阿特拉津7 d后，其叶绿素a含量显著降低，细胞结构受到破坏，种群生长受到抑制[8]；Bejarano等[9]发现，将3种常见桡足类(Enhydrosoma baruchi, Onychocamptus sp.和Paronychocamptus wilsoni)暴露于26 μg/L阿特拉津28 d后，其丰度明显降低，进而改变了底栖动物的群落结构。另有研究报道，将大西洋鲑鱼(Salmo salar)暴露于100 μg/L阿特拉津10 d后，出现了离子调节、生长和内分泌紊乱的情况[10]。鉴于阿特拉津对海洋生态系统的潜在危害，制定相应的海水水质基准(Water quality criteria, WQC)显得尤为重要。WQC指水环境中的污染物或有害因素对人体健康或水生态系统不产生有害影响的最大浓度[11]。WQC的推导方法有评估因子法(Assessment factor, AF)、物种敏感度排序法(Species sensitivity rank, SSR)和物种敏感度分布法(Species sensitivity distribution, SSD)等[12]。Chen等[13]利用AF法，将绿藻的半数效应浓度(Median effect concentration, EC50)除以100，得到阿特拉津的预测无效应浓度(Predicted no effect concentration, PNEC)为1 050 ng/L；Papadakis等[14]同样利用AF法将藻类的无观察效应浓度(No observed effect concentration, NOEC)除以10，得到阿特拉津的PNEC为10 000 ng/L。Caquet等[15]使用SSD法拟合阿特拉津对藻类的NOEC值，推导出5%物种危害浓度(Hazardous concentration for 5% of the species, HC5)值为2 200 ng/L。以上这些研究中所用的生物毒性数据较少，且主要基于单一生物类群的数据，难以为制订阿特拉津的海水WQC提供有力支撑。
为制定阿特拉津的海水WQC提供有力支撑，本研究搜集和筛选了阿特拉津对海洋生物的急性和慢性毒性数据，利用SSD法推导了阿特拉津的长期和短期海水WQC。同时，还调查了莱州湾海水中阿特拉津的污染现状，并利用本研究推导的长期海水WQC评估了阿特拉津在该海域中的生态风险。
1 材料与方法 1.1 阿特拉津阿特拉津，化学名称为2-氯-4-乙胺基-6-异丙胺基-1,3,5-三嗪，分子式为C8H14ClN5。其在海水中的半衰期约为27.7 d，主要降解产物为去乙基阿特拉津与去异丙基阿特拉津[16]。
1.2 毒性数据的搜集与筛选本研究从美国环保署(US EPA)的生态毒理数据库(https://cfpub.epa.gov/ecotox/)、Web of Science (http://apps.webofknowledge.com)和中国知网(http://www.cnki.net/)等数据库中搜集已发表的阿特拉津毒性数据，时间截至2022年1月1日。根据《HJ 1260—2022海洋生物水质基准推导技术指南(试行)》[11]和“澳大利亚和新西兰水质基准制定的指导方针”[17]对数据进行筛选。对于急性毒性数据，暴露时间以1~4 d为宜；对于慢性毒性数据，微藻的暴露时间应大于1 d，小型无脊椎动物和水生维管束植物的暴露时间应不少于7 d，大型无脊椎动物的暴露时间不少于14 d，鱼类和两栖类的幼体和成体的暴露时间应分别不少于7和21 d。搜集的毒性数据应至少涵盖3个营养级，至少包括10个物种，且涵盖以下生物类群：1种鱼科、2种甲壳科、1种非鱼类的底栖动物、1种浮游植物和1种水生维管束植物。当同一物种存在多个毒性数据时，取其几何平均值作为该物种最终的种平均急性和慢性毒性值。
1.3 水质基准推导利用搜集的毒性数据，以物种的毒性数据为x轴，以物种毒性数据的累积概率为y轴，构建SSD曲线。通过计算得到HC5值，再将HC5除以一个评估因子(取2)得到水质基准[18]。使用正态分布(Normal)、对数正态分布(Log-Normal)、对数逻辑斯蒂分布(Log-Logistic)和Bull Ⅲ分布等常用模型进行数据拟合。采用基于R语言的SSD Tools软件进行数据的拟合。
1.4 莱州湾海水样品采集如图 1和表 1所示，选取莱州湾作为调查海域，共设置了14个站位，于2021年3月进行采样。使用有机玻璃采水器采取表层(水深50 m以上)海水，通过GF/F玻璃微纤维滤膜(孔径0.22 μm，直径47 mm)进行过滤后运回实验室，以用于阿特拉津的提取。
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图 1 莱州湾取样站位 Fig. 1 Sampling stations in Laizhou Bay |
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表 1 莱州湾各站位坐标以及阿特拉津浓度 Table 1 Coordinates of sampling stations and concentrations of atrazine in the seawater |
将400 mL过滤后的表层海水与10 μL同位素内标母液(阿特拉津D5，100 μg/L)混匀后，利用预先活化的固相萃取柱(Oasis HLB)进行抽滤富集，用5 mL甲醇进行洗脱，将洗脱液收集在玻璃管中。然后将玻璃管中的洗脱液进行氮吹干燥，用复溶液(乙腈∶水=1∶9)进行复溶后，利用1 mL注射器将复溶液移出并过滤至色谱瓶中。根据Mazzella等[19]的方法，使用高效液相色谱-电喷雾质谱(HPLC-ESI-MS/MS)对阿特拉津进行测定。为验证方法的可靠性，设置了3个空白水样进行潜在污染分析，均未检测出阿特拉津。在固相萃取之前加入标准品测定回收率，监测样品预处理与仪器分析过程中的误差。本研究中阿特拉津的检出限为3 ng/L，回收率为96%，相对标准偏差为12%。
1.6 生态风险评估 1.6.1 商值法商值法的公式为
| $ R_{\mathrm{HQ}}=A_{\mathrm{EC}} / B_{\mathrm{WQC}}。$ |
式中：RHQ代表阿特拉津环境浓度与水质基准的比值；AEC代表阿特拉津在环境中的浓度；BWQC代表水质基准。根据RHQ的大小可以将风险水平分为四级：RHQ < 0.1表明阿特拉津的生态风险可以接受；0.1≤RHQ < 1表明存在低风险；1≤RHQ < 10表明存在中等风险；RHQ≥10表明存在高度风险[20-21]。
1.6.2 联合概率曲线(JPC)法本文使用风险评估软件BMC-SSD构建联合概率曲线(Joint probability curve, JPC)[22-24]。JPC是在得到毒性数据SSD曲线和阿特拉津环境浓度累积频率分布的基础上，以毒性数据SSD曲线的纵坐标为JPC的横轴，以超过毒性数值的环境浓度百分比为纵轴，形成累积剖面图，JPC与坐标轴之间的面积即为预期的生态风险[25]。
2 结果 2.1 阿特拉津的毒性数据本研究共搜集到阿特拉津对海洋生物的急性毒性数据76个，涵盖10门34科，共38个物种，如表 2所示。急性毒性数据从0.020 mg/L (扁藻(Tetraselmischuii)和鼓藻(Bellerochea polymorpha), EC50)到1 000 000 mg/L(扇蟹(Neopanopetexana), LC50) 不等。最敏感的物种与最不敏感的物种相差很大(相差约50 000倍)。在种平均急性毒性数据中，藻类和甲壳类数据最多，分别占比为44.7%和39.5%。
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表 2 阿特拉津对海洋生物的急性毒性数据 Table 2 Acute toxicity data of atrazine to marine organisms |
本研究共搜集到阿特拉津对海洋生物的慢性毒性数据32个，涵盖10门18科，共19个物种，如表 3所示。慢性毒性数据从0.004 mg/L(舟形藻(Navicula sp.), LOEC)到30.000 mg/L (月光螺(Marisa cornuarietis), NOEC) 不等。最敏感的物种与最不敏感的物种相差约6 900倍。在种平均慢性毒性数据中，藻类、甲壳动物和软体动物数据最多，分别占比为52.6%、15.8%和15.8%。
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表 3 阿特拉津对海洋生物的慢性毒性数据 Table 3 Chronic toxicity data of atrazine to marine organisms |
使用5种模型对阿特拉津的急性和慢性数据进行拟合，结果表明Anderson-Darling(AD)和Kolmogorov-Smirnov(KS)检验均为P>0.05。AIC和AICC检验结果接近，其中Log-Gumbel模型的AICC值最小，参数Delta为0，如表 4所示。因此，本研究选择Log-Gumbel模型对阿特拉津的急性和慢性毒性数据拟合SSD曲线，得到HC5值分别为13 200 ng/L (见图 2)和3 950 ng/L (见图 3)。通过HC5值除以评估因子，得到阿特拉津的短期和长期海水水生生物水质基准，分别为6 600和1 975 ng/L。
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表 4 阿特拉津海水急性和慢性毒性数据的五种拟合模型相关参数 Table 4 Five fitting model parameters for acute and chronic toxicity data of Atrazine |
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图 2 阿特拉津对海洋生物的急性毒性数据SSD拟合曲线 Fig. 2 SSD fitting curve of acute toxicity data of atrazine to marine organisms |
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图 3 阿特拉津对海洋生物的慢性毒性数据SSD拟合曲线 Fig. 3 SSD fitting curve of chronic toxicity data of atrazine to marine organisms |
在莱州湾海域的14个站位的表层海水中，阿特拉津的检出率为100%，浓度从48.15 ng/L到118.24 ng/L不等(见表 1)。从整体上看，阿特拉津的浓度呈现出由南部沿岸向北部海域递减的趋势(见图 4), 其中，站位B8和B6的阿特拉津浓度最高，分别为118.24和104.71 ng/L，站位B8的阿特拉津浓度分别是站位B3和B4的2.46和1.81倍。
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图 4 阿特拉津在莱州湾的空间分布 Fig. 4 Spatial distribution of atrazine in Laizhou Bay |
基于本文推导的长期海水水质基准(1 975 ng/L)，对阿特拉津浓度最高的站位B8进行风险评估，通过商值法得到RHQ为0.06(< 0.1)。结合莱州湾14个站位检测到的阿特拉津浓度，使用BMC-SSD软件导出联合概率曲线。如图 5所示，阿特拉津在莱州湾表层海水中的总体风险概率中值和均值分别为1.04%和0.79%。
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图 5 莱州湾表层海水中阿特拉津的联合概率曲线 Fig. 5 Joint probability curves of atrazine in surface waters of Laizhou Bay |
本研究共获得阿特拉津对38种海洋生物的平均急性毒性数据，发现绿藻门扁藻(Tetraselmis chuii)和褐藻门鼓藻(Bellerochea polymorpha)对阿特拉津最为敏感，平均急性毒性值为0.02 mg/L。相比之下，节肢动物门的扇蟹(Neopanope texana)对阿特拉津最不敏感，平均急性毒性值达到1 000 mg/L。同时，慢性毒性数据显示褐藻门舟形藻(Navicula sp.)对阿特拉津最敏感，而软体动物门的月光螺(Marisa cornuarietis)则最不敏感。有研究表明，具膜舟形藻(Navicula pelliculosa)对与阿特拉津同为三嗪类除草剂的扑草净非常敏感，慢性毒性值仅为300 ng/L[63]，这与本文得到的舟形藻对阿特拉津慢性毒性最敏感的结果一致。阿特拉津主要通过破坏植物光合系统Ⅱ (PS Ⅱ)反应中心来抑制受体和供体的电子传递[64]。浮游植物作为海洋初级生产者，通过光合作用固碳合成糖类、脂质和蛋白质等生物大分子，构成了海洋生态系统和海洋食物网初级生产力的基础[34]。在本研究中发现，浮游植物对阿特拉津最为敏感，因此特别提示应关注三嗪类除草剂对海洋初级生产力的影响。
SSD法是中国、荷兰、澳大利亚和新西兰等国家规定使用的水质基准推导方法。本研究利用5种模型拟合了阿特拉津的急性和慢性毒性数据，其中Log-Gumbel模型对数据的拟合最好，这与秦璐等[65]使用SSD法拟合除草剂高效氟吡甲禾灵的毒性数据时优选的模型一致。本研究推导阿特拉津的长期海水水质基准为1 975 ng/L，与加拿大保护水生生物的阿特拉津水质基准和欧盟规定的阿特拉津最大容许浓度(分别为1.8和2.0 μg/L)相近[5]。Moore等[66]使用SSD法拟合阿特拉津对水生自养生物的EC50值，推导出阿特拉津的HC5值为28.4 μg/L，明显高于本研究推导的水质基准。这可能是因为前者没有区分海水与淡水生物。考虑到盐度会对污染物的毒性具有明显影响，在推导海水水质基准时应该选择海洋生物毒性数据[17]。有研究表明，1 μg/L阿特拉津对大叶藻(Zostera marina L.)、丛生大叶藻(Z. caespitosa M.)和红须根虾形藻(Phyllospadix iwatensis M.)的光合作用未产生显著影响，而暴露浓度达到5 μg/L时，对3种海草的光合作用抑制率为6.89%~8.94%[67]。另有研究表明，将海葵(Exaiptasia diaphana)暴露于3 μg/L阿特拉津2周后，而其谷氨酸水平得到显著上调[68]。此外，本研究搜集到的阿特拉津对海洋生物的最低观察效应浓度(Lowest observed effect concentration, LOEC)均高于推导的阿特拉津海水水质基准。可见，本研究推导的阿特拉津海水水生生物水质基准可为海洋生物提供有效保护，因而可为制订海洋中阿特拉津的WQC提供科学依据。
本研究中，阿特拉津在莱州湾海域的浓度范围为48.15~118.24 ng/L，这与徐英江等[69]报道的浓度(6.82~83 ng/L)接近。对中国大连海域和澳大利亚大堡礁海域的多项研究发现，除草剂在近岸浓度较高，特别是河口区的阿特拉津浓度可达到0.4~3 μg/L[5, 70]。在本次阿特拉津调查中，莱州湾南部的B8站位浓度最高(118.24 ng/L)，这可能是由莱州湾南部有较多河流汇入莱州湾海域所致，如胶莱河、潍河、白浪河和小清河等，这些河流沿岸是重要的农业区，可能存在较高的阿特拉津残留。2种风险评估结果都表明，莱州湾海域的阿特拉津尚未造成明显的影响，风险可以忽略。然而，值得注意的是，莱州湾有多条河流汇入，这些河流可能携带阿特拉津。在复杂的近海河口环境中，阿特拉津浓度受季节影响较大，因此需要在时间和空间维度上进一步调查阿特拉津的污染状况。
4 结论(1) 阿特拉津的短期和长期海水水质基准分别为6 600和1 975 ng/L。
(2) 阿特拉津在莱州湾表层海水中的浓度范围为48.15~118.24 ng/L，并且在南部河口沿岸海域污染水平最高。
(3) 莱州湾海水中阿特拉津的生态风险处于可接受水平，但仍需要关注汛期入海河流排放对阿特拉津污染的影响。
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2. Shandong Provincial Key Laboratory of Restoration for Marine Ecology, Shandong Marine Resources and Environment Research Institute, Yantai 264006, China