微(纳米)塑料(micro(nano)plastics, MNPs)和抗生素是地表水体中2种新型环境污染物。环境中的微(纳米)塑料主要来源于日化用品(如合成纺织品、个人护理产品)、运输业(如合成橡胶轮胎的腐蚀)和工业生产(如塑料颗粒),通过河流运输或直接排放到海洋中[1],对水生生物造成影响。因COVID-19暴发,抗生素的市场需求暴增[2];同时,废水设施的限制,使得进入市政污水处理厂的部分抗生素随尾水排出,在自然水体中大量累积。因此,MNPs和抗生素均在水体中大量存在,其造成的长期效应会在水生生物体内积累,并通过食物链逐渐放大,甚至引起整个水生态系统的慢性毒性效应。抗生素与MNPs具有相似的来源和迁移途径,它们在水生环境中不可避免地共存,形成复合污染,因此研究抗生素与MNPs之间的相互作用及其在水生环境中的联合毒性至关重要。本文综述了MNPs和抗生素在环境中的污染现状以及两者之间的吸附方式及影响因素,概述了两者联合暴露对鱼类的毒性效应,并对两者的联合作用机制进一步的研究方向进行了展望。
1 抗生素及MNPs的污染现状(Contamination of antibiotics and MNPs)
1.1 抗生素的污染现状
抗生素具有水溶性高和易排出体外的特点,同时由于其大量使用,造成土壤和水体中抗生素的大量积累,已构成生态风险及健康威胁。尽管大多数抗生素的半衰期很短,但由于持续排放到环境中,抗生素被认为是一种“假持久性”有机污染物[3]。抗生素在医院、养殖场等特殊环境中广泛存在[4-7],已有大量文献报道了地表水体中抗生素的浓度,如表1所示。尽管抗生素的环境浓度通常处于痕量水平(ng·L-1～μg·L-1),但低浓度的抗生素仍然可能对水生生物构成风险,同时伴随着抗生素耐药菌株和抗性基因的产生和传播,对水生生态环境和人类健康造成威胁。
表1 抗生素在地表水体中的检出水平和种类
Table 1 The types and levels of antibiotics detected in surface waters
类型Types地点Locations浓度/(ng·L-1)Concentration/(ng·L-1)抗生素种类Class参考文献References河流River中国渭河西安段Xi’an Section of the Weihe River, ChinaND～270.6磺胺类、大环内酯类、喹诺酮类、四环素类Sulfonamides, macrolides, quinolones, tetracyclines[8]中国南四湖入湖河流Nansi Lake’s inflowing rivers, ChinaND～694磺胺类、大环内酯类、喹诺酮类Sulfonamides, macrolides, quinolones[9]中国渤海湾入海河流Seaborne rivers of Bohai Bay, China178.89～229.80磺胺类、喹诺酮类、四环素类Sulfonamides, quinolones, tetracyclines[10]印度亚穆纳河River Yamuna, IndiaND～19 460青霉素类、喹诺酮类、β-内酰胺类Penicillins, quinolones, β-lacams[11]湖泊Lake中国太湖Taihu Lake, ChinaND～36.472磺胺类、喹诺酮类、四环素类Sulfonamides, quinolones, tetracyclines[12]中国南四湖Nansi Lake, ChinaND～694磺胺类、大环内酯类、喹诺酮类、四环素类Sulfonamides, macrolides, quinolones, tetracyclines[9]中国东洞庭湖East Dongting Lake, ChinaND～843.49磺胺类、大环内酯类、喹诺酮类、四环素类Sulfonamides, macrolides, quinolones, tetracyclines[13]地表径流Surface runoff中国南京Nanjing, China1.958磺胺类、大环内酯类、四环素类、β-内酰胺类Sulfonamides, macrolides, tetracyclines, β-lactams[14]中国浙江Zhejiang, China508.7磺胺类、喹诺酮类、四环素类、氨霉素Sulfonamides, quinolones, tetracyclines, aminomycins[15]西班牙Spain1 300磺胺类、大环内酯类、喹诺酮类Sulfonamides, macrolides, quinolones[16]水产养殖场Aquafarm中国固城湖蟹塘Crab ponds of Lake Guchenghu, China122～1 440磺胺类、大环内酯类Sulfonamides, macrolides[17]中国江苏养殖场Aquafarm, Jiangsu, ChinaND～9 600磺胺类、四环素类Sulfonamides, tetracyclines[18]葡萄牙北部Northern Portugal2.4～10喹诺酮类、四环素类Quinolones, tetracyclines[19]污水处理厂Sewage treatment plant中国杭州Hangzhou, ChinaND～88磺胺类、大环内酯类Sulfonamides, macrolides[16]瑞典Sweden410喹诺酮类、大环内酯类Quinolones, macrolides[20]海洋Sea中国渤海湾沿海水域Coastal waters of Bohai Bay, China27.85～478.33磺胺类、喹诺酮类、四环素类Sulfonamides, quinolones, tetracyclines[10]
注:ND表示未检出。
Note: ND means not detected.
1.2 MNPs的污染现状
据报道,我国湖水中MNPs的丰度为900～34 000 个·m-3[21-25],其中长沙城市湖泊地表水中MNPs的丰度为7 050 个·m-3[24],并且超过89.5%的MNPs尺寸<2 mm,最严重的为鄱阳湖,含量达5 000～34 000 个·m-3[25]。美国南卡罗来纳州查尔斯顿港和温亚湾MNPs的检出浓度分别为(6.6±1.3) 个·m-3和(30.8±12.1) 个·m-3[26]。德国托伦斯湖中MNPs的浓度为0.14 个·m-3[27]。印度红山湖作为向钦奈市北部供水的淡水系统之一,检出MNPs的丰度为5.9 个·L-1[28]。在抗生素检出较多的区域,如污水处理厂、垃圾填埋场和蔬菜生产基地等,其周围检测出大量MNPs,丰度为4～72 个·L-1,并且粒径主要为0～50 μm,占检出颗粒的80%[29]。Wang等[30]在工业厂房、养殖场和鱼塘废水中均检测到了MNPs,其丰度分别为8～23、8～40和13～27 个·L-1,其中89%的MNPs直径<500 mm。不同来源的污水或废水之间没有明显差异,表明它们都构成了微塑料污染。生活污水处理厂的进水和出水中MNPs丰度分别为18～890 个·L-1和6～26 个·L-1,去除效率为35%～98%。南京[31]2家污水处理厂进水中的微塑料浓度分别为22.05 个·L-1和10.30 个·L-1,虽然其总去除率达到98%和97.67%,但由于日进水量巨大,因此仍有大量微塑料随尾水排放到自然水体中。在世界各地,包括偏远的极地地区,几乎都可以检测到微塑料[32-33]。因此,在不同营养层级、不同栖息地和拥有不同摄食特征的水生生物体内也发现MNPs亦不足为奇[34-35]。
MNPs和抗生素在环境中广泛分布,尤其在水体中其分布区域重叠度较大,这为它们的相互作用提供了有利条件[36]。例如,由于密集的人类活动和抗生素的大量使用,在长江口的地表水中均检测到MNPs和抗生素,它们的最高浓度分别达到10 200 个·m-3和287 ng·L-1。MNPs和抗生素也可能同时存在于北美的苏必利尔湖等沉积物中[37-38]。据报道,苏必利尔湖沉积物中的MNPs丰度为0～55 个·kg-1,同时人们也发现抗生素在苏必利尔湖的沉积物中积累[39-40]。此外,在东亚沿海循环水养殖系统和中国渤海海岸线的沉积物中,也同时检测到了MNPs和抗生素及抗性基因的存在[41-42]。因此,MNPs可以作为抗生素的载体,驱动抗生素和抗性基因在自然界中的迁移转化[43-44]。当被生物体摄入时,MNPs也会改变抗生素在生物体内的蓄积和毒性。
2 MNPs吸附抗生素(Antibiotics adsorbed by MNPs)
2.1 MNPs吸附抗生素的方式
MNPs可以通过多种物理和化学作用吸附抗生素,如范德华力、氢键、疏水相互作用和离子交换等方式,通过生物富集作用对鱼类产生危害(图1)。范秀磊等[45]认为,MNPs吸附抗生素主要经过3个阶段。第1阶段,抗生素通过疏水分配作用和范德华力吸附在MNPs表面;第2阶段,抗生素缓慢地从表面扩散到MNPs内部;第3阶段,吸附达到平衡。目前已有关于不同类型的微塑料吸附各类抗生素的相关研究,详见表2。
图1 微(纳米)塑料和抗生素联合暴露对鱼类的毒性效应
Fig. 1 The toxicity effects on fish by the joint exposure of micro(nano)plastics and antibiotics
表2 微(纳米)塑料吸附抗生素的主要方式
Table 2 The main modes for the adsorption of antibiotics by micro(nano)plastics
抗生素类型Types of antibiotics微(纳米)塑料类型Types of micro(nano)plastics吸附方式Adsorption modes参考文献References磺胺甲噁唑Sulfamethoxazole聚乳酸、聚丙烯Polylactic acid, polypropylene疏水相互作用或静电相互作用Hydrophobic interaction or electrostatic interaction[69]磺胺嘧啶Sulfadiazine聚酰胺、聚对苯二甲酸乙二醇酯、聚乙烯、聚氯乙烯、聚苯乙烯、聚丙烯Polyamide, polyethylene terephthalate, polyethylene, polyvinyl chloride, polystyrene, polypropylene范德华力和微孔填充Van der Waals forces and micropore filling[61]磺胺类抗生素Sulfonamides老化聚酰胺、聚氯乙烯Aged polyamide, polyvinyl chloride聚对苯二甲酸乙二醇酯Polyethylene terephthalate疏水相互作用Hydrophobic interaction[59]磺胺嘧啶、环丙沙星Sulfadiazine, ciprofloxacin聚乙烯Polyethylene静电相互作用Electrostatic interaction[43]环丙沙星Ciprofloxacin聚乙烯Polyethylene疏水相互作用和静电相互作用Hydrophobic interaction and electrostatic interaction[70]聚苯乙烯、聚氯乙烯Polystyrene, polyvinyl chloride疏水相互作用、π-π堆叠、静电相互作用和氢键Hydrophobic interaction, π-π stacking, electrostatic interaction and hydrogen bonds[71]诺氟沙星Norfloxacin聚苯乙烯、聚丁二酸丁二醇酯Polystyrene, polybutanediol succinate静电相互作用Electrostatic interaction[60]恩诺沙星、环丙沙星、诺氟沙星Enrofloxacin, ciprofloxacin, norfloxacin聚丙烯、聚乙烯、聚氯乙烯Polypropylene, polyethylene, polyvinyl chloride疏水相互作用和静电相互作用Hydrophobic interaction and electrostatic interaction[52]土霉素Oxytetracycin聚苯乙烯Polystyrene离子交换Ion exchange[64]四环素Tetracycline聚苯乙烯Polystyrene疏水相互作用Hydrophobic interaction[60]四环素、环丙沙星Tetracycline, ciprofloxacin聚乳酸、聚氯乙烯Polylactic acid, polyvinyl chloride氢键、π-π堆叠和静电相互作用Hydrogen bonds, π-π stacking, and electrostatic interaction[72]四环素、金霉素、土霉素Tetracycline, aureomycin,oxytetracycin聚乙烯Polyethylene范德华力和微孔填充Van der Waals forces and micropore filling[63]阿奇霉素、克拉霉素Azithromycin, clarithromycin聚乳酸、聚苯乙烯Polylactic acid, polystyrene疏水相互作用Hydrophobic interaction[73]阿莫西林Amoxicillin聚乙烯、聚对苯二甲酸乙二醇酯、聚丙烯、聚苯乙烯、聚氯乙烯Polyethylene, polyethylene terephthalate, polypropylene, polystyrene, polyvinyl chloride静电相互作用Electrostatic interaction[74]阿莫西林、四环素、环丙沙星Amoxicillin, tetracycline, ciprofloxacin聚酰胺Polyamide氢键Hydrogen bonds[46]
续表2抗生素类型Types of antibiotics微(纳米)塑料类型Types of micro(nano)plastics吸附方式Adsorption modes参考文献References磺胺嘧啶、阿莫西林、四环素、环丙沙星Sulfadiazine, amoxicillin, tetracycline, ciprofloxacin聚乙烯、聚苯乙烯、聚酰胺、聚丙烯、聚氯乙烯Polyethylene, polystyrene, polyamide, polypropylene, polyvinyl chloride氢键Hydrogen bonds[43]磺胺硫唑、磺胺美嗪、磺胺甲噁唑、环丙沙星、恩诺沙星、氧氟沙星、诺氟沙星、四环素Sulfathiazole, sulfametizine, sulfamethoxazole, ciprofloxacin, ennofloxacin, ofloxacin, norfloxacin, tetracycline聚丙烯Polypropylene氢键和疏水相互作用(原始聚丙烯)氢键和静电相互作用(老化聚丙烯)Hydrogen bonds and hydrophobic interaction (primary polypropylene)Hydrogen bonds and electrostatic interaction (aging polypropylene)[75]
在MNPs吸附抗生素的过程中,氢键的形成发挥了重要的作用。抗生素中一些特定的官能团有助于氢键的生成,例如,聚酰胺(PA)的酰胺基(质子供体基团)和阿莫西林(AMX)、四环素(TC)和环丙沙星(CIP)的羰基(质子受体基团)之间可以形成氢键[43,46]。傅里叶红外光谱分析显示的3 500 cm-1和3 100 cm-1处的峰来源于分子间氢键的相互作用,被认为是CIP、左氧氟沙星和聚苯乙烯(PS)、聚氯乙烯(PVC)之间通过氢键相连接的证据[47-48]。Yang等[49]发现随着pH值的升高,CIP上的氢离子减少,并且在PS上的吸附量降低,因此推测CIP与PS通过氢键吸附。同时,氢键也被证明是磺胺甲噁唑(SMX)在PA、PS、PVC、聚乙烯(PE)、聚丙烯(PP)和聚对苯二甲酸乙二醇酯(PET)上吸附的主要机制[50]。
大多数MNPs具有丰富的烷基基团和较强的疏水性,由疏水相互作用主导吸附过程[51]。研究表明,疏水作用在一定程度上主导了AMX、TC、CIP、甲氧苄氨嘧啶(TMP)、泰乐菌素(TYL)、SMX、磺胺甲基嗪和磺胺嘧啶在MNPs上的吸附,并且具有较高正辛醇-水分配系数(logKow)的抗生素对MNPs的亲和力更强[43,52-57]。PS和TC的结构中均具有苯环,因此二者主要通过疏水相互作用吸附在一起[58]。Fu等[59]还发现老化后的PA、PVC和PET可通过疏水相互作用吸附磺胺类抗生素。
范德华力也是MNPs吸附抗生素最常见的方式,主要由π-π和静电相互作用组成。例如PE仅通过范德华力吸附CIP、TMP和磺胺嘧啶(SDZ),并且淡水系统中CIP通过静电引力增加了在MNPs表面的吸附能力[43];当pH<6时,PS和聚丁二酸丁二醇酯(PBS)通过静电相互作用吸附诺氟沙星(NOR)[60];PA、PE、PVC、PS和PP主要通过范德华力吸附SDZ[61]。在研究对CIP、TMP和SDZ吸附机制时,研究者发现PS能够同时利用非特异性范德华力和π-π相互作用,而PE仅利用范德华力,从而导致PS对于CIP、TMP和SDZ具有较高的吸附能力[49,62]。Chen等[63]的研究也证明TC在PE上的吸附主要是由范德华力和微孔填充机制控制。
除此之外,土霉素(OTC)对PS的吸附主要由阳离子交换机制主导[64]。PE通过疏水和静电相互作用吸附CIP[65]。一些MNPs能够同时通过氢键和π-π堆积作用吸附抗生素,形成稳定结合[66]。另外,Wu等[67]提出了几个新的吸附机制。PVC上的氯原子可以作为电子受体,苯环和双酚上的羟基可以作为电子供体,从而在PVC和双酚之间形成卤素键[68]。因此,推测PVC与含羟基和苯环的抗生素之间能够形成卤素键。烷基和芳香环之间的CH/π相互作用也可以驱动PE和具有苯环的抗生素之间的吸附。由此可知,抗生素在MNPs上的吸附受到多种机制的影响。MNPs和抗生素的特定结构和性质会影响各种驱动力的贡献,导致吸附能力存在很大差异。因此,未来的研究可以集中研究吸附过程中各个机制的相对贡献。
2.2 MNPs吸附抗生素的影响因素
抗生素在MNPs上的吸附-解吸过程共同决定抗生素在MNPs上的吸附量,从而影响抗生素在环境中的迁移、分布和富集。这一吸附-解吸过程主要受到抗生素性质、MNPs类型和环境条件(如离子强度、pH和其他污染物)的影响。
2.2.1 抗生素的性质
抗生素的疏水性以及与其疏水性相关的特性(logKow和电离常数pKa)在吸附过程中发挥着重要作用[76]。郭梦函[77]研究了AMX、CIP和TC在MNPs上的吸附能力,发现吸附能力与抗生素的疏水性成正相关的关系。大多数MNPs富含烷基且疏水性较强,因此更倾向于吸附疏水性污染物[51]。Syranidou和Kalogerakis[78]的研究表明,具有较高logKow值的抗生素对MNPs的亲和力更强,因为它们具有更强的疏水性。
抗生素是可电离的化合物,其电离常数也会影响MNPs和抗生素之间的结合机制,尤其涉及到静电相互作用。抗生素的pKa、介质的pH值和MNPs的零电荷pH点(pHpzc)共同影响抗生素和MNPs之间的静电吸附过程[79]。根据抗生素的电离常数和结构,在不同的pH条件下抗生素会表现出不同的离子形态(两性离子、阳离子和阴离子),例如,在pH = 2和4时CIP主要是以阳离子形式存在[52]。Li等[43]研究了CIP在特定的pH条件下的离子形态,发现在pH为6.7～7.1时,CIP以两性离子、阴离子和阳离子的形式存在;而在pH为8.0时,CIP的主要存在形式为两性离子和阴离子。在这2种情况下,MNPs的pHpzc均小于环境pH值,呈现负电性。因此在pH为8.0时,MNPs和CIP之间的静电排斥作用增强,从而降低CIP的吸附水平。TC的pKa2 = 7.7,当pH<7.7时,TC主要以两性离子和阳离子形式存在,带负电的MNPs可通过静电作用吸附TC;当pH>7.7时,TC的主要存在形式为阴离子,由于与MNPs静电排斥,此时TC的吸附量大大降低[80]。
除此之外,抗生素的极性也对MNPs-抗生素的吸附过程有影响,具有多个极性官能团的抗生素可以促进MNPs的吸附。例如,喹诺酮类抗生素具有较多的极性官能团如羧基、羟基等,易于与环境中的MNPs发生吸附作用;磺胺类抗生素仅有苯氨基和酰胺基,因此MNPs对其的吸附能力较弱[81]。
2.2.2 MNPs的性质
MNPs由于其具有比表面积大、疏水性强和流动性高的特点,在环境中可以积聚各种毒素和化学污染物,并作为远距离运输污染物的载体。MNPs的性质,如极性、比表面积和结晶度等,对污染物的吸附能力有很大影响[82-83]。
MNPs的官能团和极性在MNPs-抗生素的吸附过程中起主导作用。PS、PP和PE通常是非极性塑料,而PVC、PET和PA是极性的。例如,强极性聚合物PA对磺胺甲噁唑和磺胺甲嗪具有比PE更强的吸附能力,这是由于PA中的酰胺基团(质子供体基团)和抗生素结构中存在的羰基(质子受体基团)之间形成了氢键,从而增强了吸附作用[84]。同样地,Fu等[85]也发现由于形成了稳定的氢键,PA对SDZ、AMX、TC、CIP、TMP的吸附能力超过了PS。
MNPs的比表面积越大,意味着吸附位点越多,因此可以吸附的污染物的量就越大。Li等[43]发现PS、PP,尤其是PA的孔隙结构较为发达,使得这3种MNPs对AMX、TC和CIP的吸附能力高于PE和PVC。其次,对于特定类型的MNPs,尺寸较小的MNPs通常具有较大的比表面积,从而对TC具有较高的吸附能力。然而,MNPs的粒径并不总是与其比表面积成反比。PS (50 nm)的实测比表面积(63.4 m2·g-1)低于理论值114.3 m2·g-1,这可能是由于PS的团聚导致比表面积降低[86]。粒径大小仅在一定范围内影响MNPs的吸附能力,最终吸附的效果取决于比表面积。
MNPs另一个影响吸附过程的特性是结晶度。结晶度是聚合物中结晶区域所占的比例,MNPs具有无定形和结晶区域,无定形区域由不规则排列的长链组成,结晶区域则由规则排列成几何晶格的链段构成[87-88]。有机污染物对无定形区域的亲和性大于结晶区域,说明MNPs的结晶度越低,其对有机污染物的吸附能力越强。Liu等[48]发现低结晶度PVC对CIP的吸附能力显著高于PS。Gong等[89]研究报道在玻璃化转变温度下,聚丁二酸丁二醇酯(PBS)表现为橡胶状聚合物,而聚乳酸(PLA)表现为玻璃状聚合物。玻璃状聚合物PLA的分子链密集且交联,阻碍了有机污染物的移动。因此,PBS的吸附能力大于PLA。然而,关于MNPs的结晶度在抗生素吸附中的作用仍缺乏明确的结论。
此外,老化作用也能够增加MNPs的吸附能力。其原因有多个方面,老化后MNPs颗粒碎裂从而导致比表面积增加,其表面含氧官能团增加也会导致MNPs极性和表面性质发生变化,以及MNPs的表面形成生物膜,可通过降低其疏水性来增强MNPs的吸附能力[48,90]。Yu等[91]在我国东山湾的海鱼养殖场对5种微塑料进行原位老化,并探究其在海洋养殖环境中对抗生素的吸附情况及影响因素。通过现场的原位吸附实验,研究发现与PS、PP、PE、PVC和PET相比,老化PP由于具有更发达的孔隙结构,因而对抗生素具有更高的吸附效率。
2.2.3 环境因素
离子强度在MNPs吸附抗生素的过程中发挥着重要的作用,其作用机理是通过影响双层膜的厚度和界面电位来控制吸附剂与吸附质表面之间的静电和非静电相互作用,从而影响两者的结合[92]。NaCl是水环境中最常见的离子,随着NaCl浓度的升高,钠离子取代了MNPs表面酸性官能团中的氢离子,抑制了氢键的形成,使得MNPs对抗生素的吸附量显著下降[93-95]。然而一些阳离子,如Cr3+、Zn2+,与抗生素在MNPs表面具有金属桥连作用,反而促进了吸附作用[47,96]。
pH值在一定程度上也决定了静电相互作用,从而影响抗生素在MNPs上的吸附。例如,Xue等[97]发现当pH值大于磺胺类抗生素的等电点时,一部分磺胺带负电,此时与带负电的MNPs之间存在静电排斥,导致吸附能力下降。Wang等[83]发现在低pH条件下,阴离子形式的全氟辛烷磺酸(PFOS)在PE上的吸附能力高于非离子形态的PFOS,并且PFOS在MNPs上的吸附量随着pH的升高而降低。在淡水和海水中,MNPs对抗生素的吸附能力也不尽相同。在pH范围为6.7～7.1的淡水中,抗生素的存在形式大多为两性离子和阴离子,但仍存在部分阳离子形式;海水的pH值比淡水更高,抗生素几乎都以两性离子和阴离子的形式存在,与携带负电荷的MNPs具有较高的静电斥力。因此与淡水相比,MNPs在海水中吸附抗生素的程度较低。Yang等[49]认为,PS在海水中对CIP的吸附能力远低于在去离子水中的吸附能力,可能是因为海水中存在的许多离子加速了溶液环境中电子的流动。
当水体中存在其他有机污染物和重金属时,会影响MNPs对抗生素的吸附。相较于抗生素,重金属离子通过离子交换更容易吸附到MNPs上[98-100]。因此当重金属与抗生素共存时,MNPs的大多数活性吸附位点被重金属离子填充,从而导致其吸附抗生素的能力受到限制。同时,重金属和MNPs之间的静电相互作用也会导致重金属被吸附,占据MNPs表面的吸附位置,从而降低MNPs对抗生素的吸附[101]。然而,多价金属离子又可以和抗生素的负电荷络合,在抗生素、金属离子和MNPs之间形成共价键,从而提高抗生素的吸附量[102]。
由于表面活性剂会改变污染物的界面特性,因此当水体中存在表面活性剂时会影响MNPs对抗生素的吸附能力。例如,十二烷基苯磺酸钠(SDBS)能与PS和PE结合,提高PS和PE的表面电负性并且降低比表面积和孔隙率,使其在保持基本晶体结构的同时表面官能团略有改变,大大提高了PS和PE对OTC和NOR的饱和吸附率。SDBS也会增强MNPs的亲水性,使其更易于吸附溶解在水中的抗生素[97]。腐殖质也会影响抗生素在MNPs上的吸附,例如,郭梦函[77]研究了3种抗生素(AMX、CIP和TC)在4种MNPs(PVC、PS、PP和PE)上的吸附情况,发现在不同浓度腐殖酸的条件下,3种抗生素在MNPs上的吸附能力先降低后升高。低浓度下3种抗生素在MNPs上的吸附量均呈现下降的趋势,这是因为腐殖质与抗生素竞争MNPs上的吸附位点,导致吸附量降低。而高浓度下腐殖质吸附到MNPs上会形成包裹层,与抗生素发生阳离子π键和π-π给体受体作用,从而导致抗生素在MNPs上的吸附量增加。此外,当水体中存在蛋白质时,蛋白质会在老化MNPs上形成蛋白质电晕,从而加强对磺胺的吸附能力[59]。
3 MNPs与抗生素联合暴露对鱼类的毒性(The joint toxicity of MNPs and antibiotics to fish)
水体环境中大量的MNPs和抗生素会对水生生物产生毒性影响,造成严重的生态风险,并且还可能通过食物链转移和富集,从而威胁人类健康[103-105]。大量文献显示MNPs-抗生素的复合污染可能引起生物体分子组织、细胞和行为方面的改变,导致生物体的损伤[106],因此MNPs和抗生素联合毒性的研究势在必行。鱼类因其容易获得、实验室易饲养并且对毒物敏感的特征,常被用于水环境中污染物毒性的研究。已有大量文献报道了各类抗生素对斑马鱼的毒性,如表3所示。在生物体内,MNPs通过吸附抗生素并干扰代谢来增强抗生素的生物积累,但不能被水生生物摄取的大尺寸MNPs会降低抗生素的生物积累。由于抗生素在外部环境和生物体内与MNPs的吸附以及它们对同一生物靶标的作用,MNPs增强/缓解了抗生素对生物体的毒性。MNPs和抗生素对鱼类的联合毒性主要表现在肠道、肝脏、神经、生殖和发育毒性等方面。
表3 抗生素对斑马鱼的毒性
Table 3 The toxicity effects of antibiotics on zebrafish
抗生素类型Class抗生素种类Types暴露方式Exposure对斑马鱼的毒性Toxicity to zebrafish参考文献References四环素类Tetracyclines土霉素Oxytetracycline金霉素Chlorotetracycline四环素Tetracycline急性暴露Acute exposure胚胎孵化延迟,诱导氧化应激Embryo hatching was delayed and oxidative stress was caused[107]慢性暴露Chronic exposure探索行为、多动和冻结行为增加Exploration behavior, hyperactivity and freezing behavior were increased导致炎症反应,扰乱肠道菌群Inflammatory response was induced and gut flora were disrupted[108][108-110]慢性暴露Chronic exposure运动距离减少,认知行为下降,攻击行为增加Motor distance was decreased, cognitive behavior was decreased, and aggressive behavior was increased[111]急性暴露Acute exposure胚胎孵化延迟,体长变短,卵黄囊肿,游囊发育受阻,引起氧化应激及细胞凋亡Embryo hatching was delayed, body length became shorter, yolk cyst appeared, follicle development was blocked, and oxidative stress and cell apoptosis were induced[112]慢性暴露Chronic exposure肝脏脂质空泡化,肝脏代谢途径失调Liver lipid vacuolation and liver metabolic pathway disorders[113]β-内酰胺类β-lactams阿莫西林Amoxicillin头孢噻肟钠Cefotaxime sodium阿莫西林Amoxicillin头孢他啶Ceftazidime急性暴露Acute exposure慢性暴露Chronic exposure胚胎过早孵化Premature hatching of embryos[107]幼鱼体长变短Larvae became shorter in length[114]社交行为减少,引起氧化应激Social behavior was reduced, and oxidative stress was caused[115]运动距离增加,攻击行为加剧Movement distance was increased and aggressive behavior was intensified[111]喹诺酮类Quinolones恩诺沙星Enrofloxacin诺氟沙星Norfloxacin加替沙星Gatifloxacin环丙沙星Ciprofloxacin诺氟沙星Norfloxacin环丙沙星Ciprofloxacin急性暴露Acute exposure慢性暴露Chronic exposure体长缩短、脊柱弯曲、心包水肿和卵黄囊肿Shortening of body length, curvature of spine, pericardial edema and yolk cyst[113-114]抑制胚胎孵化率,提高死亡率和畸形率Embryo hatching rate was inhibited, and mortality and deformity rate were increased[116]心包水肿,心率和心输出下降Pericardial edema and decreased heart rate and cardiac output[117]心率和心输出下降Heart rate and cardiac output were decreased[117]引起氧化应激,降低亲代的产卵率Oxidative stress was induced and oviposition rate of the maternal generation was reduced[118]运动距离减少,认知行为下降,攻击行为增加Movement distance was decreased, cognitive behavior was decreased, and aggressive behavior was increased[111]
续表3抗生素类型Class抗生素种类Types暴露方式Exposure对斑马鱼的毒性Toxicity to zebrafish参考文献References大环内酯类Macrolides红霉素Erythromycin阿奇霉素、克拉霉素、替米考星、泰利霉素Azithromycin, clarithromycin,timicacin, telithromycin急性暴露Acute exposure胚胎孵化延迟,心率上升Embryo hatching was delayed and heart rate was increased[119]肝脏退化,肝脏大小改变,肝脏脂肪变性Liver involution, changes in liver size, and hepatic steatosis[120-121]心率低浓度时上升,高浓度时下降Heart rate was increased at low concentrations and decreased at high concentrations[122]氨基糖苷类Aminoglycosides庆大霉素、新霉素Gentamicin, neomycin庆大霉素Gentamicin急性暴露Acute exposure侧线毛细胞脱落,减少运动行为Lateral line hair cells were shed and motor behavior was reduced[123-125]急性肾衰竭,出现水肿Acute renal failure and edema formation[126]磺胺类Sulfonamides磺胺甲噁唑Sulfamethylthiazole急性暴露Acute exposure胚胎孵化延迟,体长变短,引起氧化应激和炎症Embryo hatching was delayed, body length became shorter, and oxidative stress and inflammation were caused[127-128]慢性暴露Chronic exposure引起氧化应激,降低亲代产卵量,子代孵化率和存活率降低Oxidative stress was caused, the amount of parental egg production was reduced, and the hatching and survival rate of the offspring were decreased[108,118]
3.1 对肠道的影响
肠道是鱼类重要的消化和营养获取器官。鱼类是较低等的脊椎动物,消化能力弱,肠道干细胞分化成更多功能性细胞(如杯状细胞、淋巴细胞和柱状上皮细胞)来吸收营养或分泌消化液以应对外部刺激[129],并且肠道微生物群的干扰可导致宿主的生理功能障碍和某些疾病[130]。因此肠道受损会引起鱼类多种疾病的发生,增加健康受损的风险。
Liu等[131]发现PS (80 nm)加剧了TC引起的对草鱼(Ctenopharyngodon idella)幼鱼的氧化应激损伤,并且伴有结肠腺癌的风险。PS和TC联合暴露组的幼鱼出现小肠绒毛受损和萎缩等组织损伤症状,鳃细胞也出现组织形态学改变,包括毛细血管扩张、上皮板层提升和板层融合。Zhang等[132]的研究结果与Liu等[131]的相似,对照组鲫鱼(Carassius auratus)的肠绒毛是有序的,罗红霉素(ROX)暴露导致肠上皮细胞空泡化和肠道纤毛受损,而在老化MNPs (5 μm和50 μm)和ROX联合暴露组中,肠细胞空泡化、纤毛受损和淋巴细胞浸润的现象更加严重。老化MNPs的载体效应因其官能团和比表面积的变化而增强,能够增加ROX在鲫鱼体内的累积。与对照组相比,单独暴露ROX、PS和老化PS导致脂肪酶(LPS)和淀粉酶(AMS)活性显著增加,而联合暴露组的LPS和AMS活性显著降低。并且ROX和老化PS联合暴露组比ROX和PS联合暴露组诱导更多的LPS和AMS抑制。与对照组相比,ROX和PS单独暴露提高了肠道中抗性基因(ARGs)的丰度,而老化PS组并没有明显的差异,但是ROX和PS/老化PS的联合暴露显著提高了肠道中ARGs的丰度,尤其ROX和老化PS联合暴露组,使得ARGs增加了69.8%[133]。因此,可推断PS/老化PS与ROX的联合作用可能表现为协同作用。有研究表明,MNPs可充当污染物的载体使其表面更有利于抗生素和ARGs的富集[134],并且肠道表面活性剂有利于提高MNPs吸附污染物的解吸速率[135],因此PS和ROS的协同作用可能是因为携带抗生素的MNPs在鱼类肠道中解吸释放抗生素,被生物体吸收,从而增强了毒性。据报道,磺胺二甲嘧啶(SMZ)会对青鳉(Oryzias melastigma)肠道菌群的稳定和抗氧化性能产生显著影响,而PS (100 nm)仅轻微改变肠道菌群的组成。然而将PS与SMZ联合使用却能够减轻SMZ的整体肠道毒性。由此猜测PS与SMZ之间的联合作用可能为拮抗关系,从而减轻了SMZ对青鳉鱼的肠道毒性[136]。Liao等[137]发现PS (100 nm)和TC的联合暴露导致海洋青鳉肠道微生物网络的关键类群发生变化,与任何单一暴露组相比,联合暴露组更显著降低了肠道微生物网络的复杂性和稳定性,为外源微生物的入侵和定植提供了机会,表明PS和TC对肠道微生物组具有协同作用。
3.2 对肝脏的影响
从生理学上讲,鱼类与哺乳动物的肝脏具有相同的基本代谢功能,包括营养物质的加工和储存、酶和其他辅助因子的合成、胆汁的形成和排泄以及外源化合物的代谢,并且有助于防止血液在鱼体内凝固。
Zhang等[132]研究发现老化PS (500 nm、5 μm和50 μm)加剧了ROX对鲫鱼肝脏的损伤,在联合暴露组的肝脏组织中观察到肝细胞空泡化和肝细胞簇排列,并且显示出更高的氧化应激水平,其中500 nm老化PS与ROX共同暴露显著加重了肝脏的嗜酸性坏死。Liu等[131]也有相同的实验结果,PS (80 nm)和TC联合暴露显著提高了草鱼肝脏的总抗氧化能力(T-AOC)、超氧化物歧化酶(SOD)和过氧化氢酶(CAT)的水平,表明PS可能增强TC引起的氧化损伤。系统发育树显示MMP2和MMP9在肝细胞癌中相对保守,而MMP2的mRNA表达与TGFβ1、IL-8和MMP9的水平呈现显著的正相关性。PS和TC的联合暴露显著上调了这些基因,因此联合暴露会提高草鱼幼鱼患肝细胞癌的风险。Zhang等[138]的研究报道PS (100 nm)能够增加ROX在罗非鱼(Oreochromis niloticus)肝脏中的生物富集,当暴露于ROX (50 μg·L-1)时,其在肝脏中的蓄积量达到了(4 307.1±186.5) μg·kg-1。与单独暴露ROX相比,暴露于ROX和PS的罗非鱼鱼肝中细胞色素P450酶的活性显著降低,表明PS的存在可能影响ROX的肝脏代谢。然而,研究者也发现ROX和PS联合暴露导致SOD活性显著增加,丙二醛(MDA)含量降低,表明ROX和PS联合暴露下的鱼肝氧化损伤得到减轻。因此,MNPs和抗生素的联合暴露可能会引起鱼类的复杂反应。在这方面,需要进一步的研究来验证鱼类的长期生化反应。此外,当MNPs或抗生素进入鱼类肠道后,肠道微生物群失衡,会导致鱼类脂质代谢紊乱,诱导肝脏发生炎症。研究报道暴露于PS和TC引起的肠道微生物组失衡,使得肠道微生物网络中的关键物种与肝脏代谢指标之间的关联减少,可能间接影响肝脏代谢[137]。
3.3 神经毒性
鱼类具有高度发达的神经系统,在生命活动中起到重要的调控作用。鱼类的情绪、感觉和学习行为是由大脑控制的,这些区域虽在解剖学上不同,但在功能上与哺乳动物非常相似。在鱼类的中枢神经系统中,神经调节剂(包括多巴胺、去甲肾上腺素、血清素、组胺和乙酰胆碱)在鱼类的运动控制和感觉处理中发挥关键作用,还参与更具体的行为调节[139]。目前对于环境污染物的神经毒性效应研究主要集中在其对行为的影响以及对生物电化学信号转导的干扰,尤其是神经递质的干扰。在各种神经递质中,乙酰胆碱酯酶(AChE)是生物神经传导中的一种关键性酶,提供了潜在神经肌肉胆碱能破坏的信息,常被作为评估环境污染物神经毒性的主要指标[140]。
许多研究表明,无论是单独暴露抗生素还是与MNPs联合暴露,都会显著抑制AChE的活性,但联合暴露组是否比抗生素暴露组产生更高的抑制效果的结果却不一致。例如,ROX和老化MNPs (500 nm)联合暴露比ROX单独暴露造成更强的AChE抑制[132],但是Zhang等[138]研究发现ROX与不同浓度的PS (100 nm)联合暴露后的AChE活性均高于ROX组,并且呈现浓度依赖性。Huang等[141]使用老化PS与SMX联合暴露也得到了类似的结果,与单独使用SMX相比,老化PS和SMX联合暴露组的AChE活性显著增加,因此老化PS减轻了SMX诱导的神经毒性。一种可能的解释是吸附在MNPs表面的抗生素因不能与突触间隙中的AChE直接接触,从而减轻了抗生素的抑制作用。对于抗生素和MNPs联合暴露对鱼类的行为影响研究较少,Lu等[142]发现,SMX暴露降低了幼鱼的游泳频率,而(327.3±72.1) nm PS与SMX联合暴露加剧了这一现象。此外,MNPs能够显著增加ROX在罗非鱼脑组织中的蓄积,经过ROX (50 μg·L-1)和PS (100 μg·L-1)联合暴露14 d后,其脑组织ROX的蓄积量达到(2 907.5±225.0) μg·kg-1。同时ROX暴露会导致成鱼脑组织出现炎症和脊髓内水肿的现象,ROX和老化MNPs (500 nm)的共同暴露则会诱导更严重的感染性炎症和脊髓内水肿,并且出现毛细胞星形细胞瘤[132]。
3.4 生殖毒性
鱼类具有许多使其容易受到毒性影响的特征,包括产卵、体外受精、胚胎在母体外发育、富含脂质的卵有利于疏水性外源物质的积累,以及环境因素对性别分化和繁殖的显著影响[143]。最近的一项研究表明,斑马鱼母体接触抗生素混合物(100 μg·L-1)会导致F1代幼鱼肠道疾病和生长迟缓[144]。而长期接触MNPs (10 mg·L-1)的斑马鱼(Danio rerio)和青鳉鱼,其产卵率分别下降70%和42%,并且斑马鱼F1代幼鱼出现过度活跃的现象[145]。这些结果表明,抗生素和MNPs可能会对鱼类产生跨代的不利影响。
对于2种及2种以上相互作用的污染物对鱼类生殖毒性的报道较少。He等[146]发现F0代青鳉接触PS (100 nm)和SMZ会导致F1代体质量下降,并且会导致孵化后2个月的雄性F1代肠道微生物群的组成发生改变。然而,接触SMZ和PS混合物的F1代比单独接触PS的F1代雄性体质量下降趋势有所减少。RT-qPCR结果显示SMZ和PS联合组的肝脏igf1表达水平显著高于PS组,并且宿主能量代谢相关类杆菌门的相对丰度也更高,表明亲本暴露于SMZ和PS的混合物可能以不同于PS的方式影响F1代稻田鱼幼鱼肠道的微生物定植,并且以协同的方式增强毒性效应。
3.5 发育毒性
最近的研究发现,暴露于MNPs会导致幼鱼孵化率降低,胚胎存活率降低以及发育过程中的能量代谢紊乱。在发育阶段,粒径较小的纳米塑料对其存活率有着更严重的影响。此外,暴露于MNPs会通过形成病理性血管,诱导尾静脉丛的异常发育,从而导致斑马鱼出现生长退化,尾部体长缩短和尾部组织受损等现象[147]。同时,抗生素对鱼类发育毒性的影响主要表现在胚胎畸形、孵化率降低、氧化应激水平增加、心率变化及游泳行为的影响[148]。
Lu等[142]收集了斑马鱼发育的多个身体区域和阶段的数据,发现暴露于PS和SMZ会导致高死亡率(25.0±7.5)%和高畸形率(20%～35%)的发生。其生理毒性表现为抽动显著减少(31.1%～37.0%)、游泳频率减少(26.9%～36.8%)和心率显著增加(19.0%～20.9%)。同时PS和SMZ联合暴露也导致了卵黄蛋白原、17β-雌二醇、睾酮和三碘甲状腺原氨酸水平的上升,对斑马鱼产生内分泌干扰效应。根据发育毒性、生理毒性和内分泌毒性的评估结果,PS和SMZ联合暴露对斑马鱼的毒性比PS和SMZ单独暴露更强,因此我们猜测PS和SMZ之间存在协同作用。Fonte等[149]研究发现头孢氨苄(CEF)单独暴露以及与PE (1～5 μm)联合暴露虾虎鱼(Pomatoschistus microps)幼鱼时的毒性曲线显著不同,其96 h的半致畸浓度值(EC50)分别为3.8 mg·L-1和5.2 mg·L-1,表明在PE存在下,CEF对虾虎鱼幼鱼的毒性降低。同时CEF和PE联合暴露也导致了幼鱼捕食能力的下降和脂质过氧化水平的升高,但是改变程度比单独暴露CEF或PE时低,我们推测CEF和PE之间可能存在拮抗作用。
4 总结与展望(Summary and outlook)
由于抗生素和MNPs在环境中的广泛分布,其对水生生物联合毒性效应的研究势在必行。根据已有的研究,抗生素和MNPs通过范德华力、氢键、离子交换、π-π相互作用和静电作用等方式相结合,暴露于水体中会导致鱼类的功能障碍,并诱发肠道、肝脏、神经以及生殖系统的相关疾病,而抗生素和MNPs对鱼类毒性的相互作用机制还未深入研究。MNPs大多数是疏水性的,而抗生素多为亲水性。我们猜测拮抗作用可能是因为MNPs吸附抗生素后一方面降低了污染物的浓度,导致生物可利用性降低,另一方面MNPs吸附抗生素后颗粒粒径变大,被生物体当作异物排出了体外,从而降低了抗生素对水生生物的毒性。而MNPs和抗生素的协同作用可能是由于水生生物摄入吸附抗生素的MNPs时,MNPs通过解吸作用将抗生素转运到生物体中,导致抗生素在生物体内积累,对肝脏等器官造成损伤,从而发挥协同效应。
然而,鉴于MNPs、抗生素和环境条件的复杂性,MNPs和抗生素的联合毒性机制可能更为复杂。抗生素和MNPs的吸附行为受到抗生素的种类、MNPs的物理化学性质和添加剂、环境因素(pH和离子强度)以及共存环境中的有机污染物、重金属、表面活性剂和生物大分子的影响,从而进一步影响其对水生生物的联合毒性效应。因此,一方面需要吸附-解吸附动力学的研究,以揭示MNPs和抗生素在水体环境和生物体中的吸附行为。另一方面,需要借助分子生物学技术深入研究MNPs-抗生素作用于水生生物的关键基因和通路,探索其联合毒性效应机制。此外,还需要在与环境相关的条件下进行原位或体内实验,进一步研究抗生素和MNPs联合作用对环境和水生生物健康的影响。
[1] Wang C H, Zhao J, Xing B S. Environmental source, fate, and toxicity of microplastics [J]. Journal of Hazardous Materials, 2021, 407: 124357
[2] Rizvi S G, Ahammad S Z. COVID-19 and antimicrobial resistance: A cross-study [J]. The Science of the Total Environment, 2022, 807(Pt 2): 150873
[3] Carvalho I T, Santos L. Antibiotics in the aquatic environments: A review of the European scenario [J]. Environment International, 2016, 94: 736-757
[4] 周瑞飞, 王少坡, 常晶, 等. 水中抗生素污染现状及高级氧化技术研究进展[J]. 天津城建大学学报, 2021, 27(5): 356-362
Zhou R F, Wang S P, Chang J, et al. Pollution status of antibiotics in water and research progress of advanced oxidation technology [J]. Journal of Tianjin Chengjian University, 2021, 27(5): 356-362 (in Chinese)
[5] 赵亚奇, 温沁雪, 杨莲, 等. 废水中磺胺甲噁唑在A/O-MBR工艺中的去除机理[J]. 中国给水排水, 2017, 33(17): 26-31
Zhao Y Q, Wen Q X, Yang L, et al. Removal principal of sulfamethoxazole in wastewater treatment in A/O-MBR [J]. China Water &Wastewater, 2017, 33(17): 26-31 (in Chinese)
[6] 于明琪, 钟传青, 周英萍, 等. 乡镇医院污水环境中抗生素耐药细菌的分离与分析[J]. 环境生态学, 2021, 3(6): 74-78
Yu M Q, Zhong C Q, Zhou Y P, et al. Isolation and analysis of antibiotic-resistant bacteria in sewage environment of township hospitals [J]. Environmental Ecology, 2021, 3(6): 74-78 (in Chinese)
[7] 林茂宏, 王震, 张开明, 等. 电氧化耦合陶瓷膜处理南方某医院污水的研究[J]. 水处理技术, 2022, 48(1): 112-117
Lin M H, Wang Z, Zhang K M, et al. Research on electro-oxidation coupled with ceramic membrane for treatment of sewage in a hospital located in South China [J]. Technology of Water Treatment, 2022, 48(1): 112-117 (in Chinese)
[8] Wang J W, Wei H, Zhou X D, et al. Occurrence and risk assessment of antibiotics in the Xi’an section of the Weihe River, northwestern China [J]. Marine Pollution Bulletin, 2019, 146: 794-800
[9] Zhang G D, Liu X H, Lu S Y, et al. Occurrence of typical antibiotics in Nansi Lake’sinflowing rivers and antibiotic source contribution to Nansi Lake based on principal component analysis-multiple linear regression model [J]. Chemosphere, 2020, 242: 125269
[10] Han Q F, Song C, Sun X, et al. Spatiotemporal distribution, source apportionment and combined pollution of antibiotics in natural waters adjacent to mariculture areas in the Laizhou Bay, Bohai Sea [J]. Chemosphere, 2021, 279: 130381
[11] Mutiyar P K, Mittal A K. Occurrences and fate of selected human antibiotics in influents and effluents of sewage treatment plant and effluent-receiving River Yamuna in Delhi (India) [J]. Environmental Monitoring and Assessment, 2014, 186(1): 541-557
[12] Xu Z A, Li T, Bi J, et al. Spatiotemporal heterogeneity of antibiotic pollution and ecological risk assessment in Taihu Lake Basin, China [J]. The Science of the Total Environment, 2018, 643: 12-20
[13] Guo X C, Song R R, Lu S Y, et al. Multi-media occurrence of antibiotics and antibiotic resistance genes in East Dongting Lake [J]. Frontiers in Environmental Science, 2022, 10: 866332
[14] Guo X Y, Xiaojun L, Zhang A G, et al. Antibiotic contamination in a typical water-rich city in southeast China: A concern for drinking water resource safety [J]. Journal of Environmental Science and Health Part B, Pesticides, Food Contaminants, and Agricultural Wastes, 2020, 55(3): 193-209
[15] Lei Y Y, Li F F, Ouyang J, et al. Environmental distribution characteristics and source analysis of antibiotics in Zhejiang area [J]. Progress in Chemistry, 2021, 33(8): 1414-1425
[16] Danner M C, Robertson A, Behrends V, et al. Antibiotic pollution in surface fresh waters: Occurrence and effects [J]. Science of the Total Environment, 2019, 664: 793-804
[17] Wang W X, Gu X H, Zhou L J, et al. Antibiotics in crab ponds of Lake Guchenghu Basin, China: Occurrence, temporal variations, and ecological risks [J]. International Journal of Environmental Research and Public Health, 2018, 15(3): 548
[18] Wei R C, Ge F, Huang S Y, et al. Occurrence of veterinary antibiotics in animal wastewater and surface water around farms in Jiangsu Province, China [J]. Chemosphere, 2011, 82(10): 1408-1414
[19] Pereira A M, Silva L J, Meisel L M, et al. Fluoroquinolones and tetracycline antibiotics in a Portuguese aquaculture system and aquatic surroundings: Occurrence and environmental impact [J]. Journal of Toxicology and Environmental Health Part A, 2015, 78(15): 959-975
[20] Östman M, Lindberg R H, Fick J, et al. Screening of biocides, metals and antibiotics in Swedish sewage sludge and wastewater [J]. Water Research, 2017, 115: 318-328
[21] Mao R F, Hu Y Y, Zhang S Y, et al. Microplastics in the surface water of Wuliangsuhai Lake, Northern China [J]. The Science of the Total Environment, 2020, 723: 137820
[22] Wang W F, Ndungu A W, Li Z, et al. Microplastics pollution in inland freshwaters of China: A case study in urban surface waters of Wuhan, China [J]. Science of the Total Environment, 2017, 575: 1369-1374
[23] Wang W F, Yuan W K, Chen Y L, et al. Microplastics in surface waters of Dongting Lake and Hong Lake, China [J]. The Science of the Total Environment, 2018, 633: 539-545
[24] Yin L S, Jiang C B, Wen X F, et al. Microplastic pollution in surface water of urban lakes in Changsha, China [J]. International Journal of Environmental Research and Public Health, 2019, 16(9): 1650
[25] Yuan W K, Liu X N, Wang W F, et al. Microplastic abundance, distribution and composition in water, sediments, and wild fish from Poyang Lake, China [J]. Ecotoxicology and Environmental Safety, 2019, 170: 180-187
[26] Gray A D, Wertz H, Leads R R, et al. Microplastic in two South Carolina Estuaries: Occurrence, distribution, and composition [J]. Marine Pollution Bulletin, 2018, 128: 223-233
[27] Tamminga M, Stoewer S C, Fischer E K. On the representativeness of pump water samples versus manta sampling in microplastic analysis [J]. Environmental Pollution, 2019, 254(Pt A): 112970
[28] Gopinath K, Seshachalam S, Neelavannan K, et al. Quantification of microplastic in red hills lake of Chennai City, Tamil Nadu, India [J]. Environmental Science and Pollution Research, 2020, 27(26): 33297-33306
[29] Shi J Y, Dong Y B, Shi Y Y, et al. Groundwater antibiotics and microplastics in a drinking-water source area, Northern China: Occurrence, spatial distribution, risk assessment, and correlation [J]. Environmental Research, 2022, 210: 112855
[30] Wang F, Wang B, Duan L, et al. Occurrence and distribution of microplastics in domestic, industrial, agricultural and aquacultural wastewater sources: A case study in Changzhou, China [J]. Water Research, 2020, 182: 115956
[31] Yuan F, Zhao H, Sun H B, et al. Abundance, morphology, and removal efficiency of microplastics in two wastewater treatment plants in Nanjing, China [J]. Environmental Science and Pollution Research International, 2021, 28(8): 9327-9337
[32] Morgana S, Ghigliotti L, Estévez-Calvar N, et al. Microplastics in the Arctic: A case study with sub-surface water and fish samples off Northeast Greenland [J]. Environmental Pollution, 2018, 242(Pt B): 1078-1086
[33] Kanhai D K, Gårdfeldt K, Lyashevska O, et al. Microplastics in sub-surface waters of the Arctic Central Basin [J]. Marine Pollution Bulletin, 2018, 130: 8-18
[34] Bour A, Avio C G, Gorbi S, et al. Presence of microplastics in benthic and epibenthic organisms: Influence of habitat, feeding mode and trophic level [J]. Environmental Pollution, 2018, 243(Pt B): 1217-1225
[35] Digka N, Tsangaris C, Torre M, et al. Microplastics in mussels and fish from the northern Ionian Sea [J]. Marine Pollution Bulletin, 2018, 135: 30-40
[36] Wang Y H, Yang Y N, Liu X, et al. Interaction of microplastics with antibiotics in aquatic environment: Distribution, adsorption, and toxicity [J]. Environmental Science &Technology, 2021, 55(23): 15579-15595
[37] Yan C X, Yang Y, Zhou J L, et al. Antibiotics in the surface water of the Yangtze Estuary: Occurrence, distribution and risk assessment [J]. Environmental Pollution, 2013, 175: 22-29
[38] Zhao S Y, Zhu L X, Wang T, et al. Suspended microplastics in the surface water of the Yangtze Estuary System, China: First observations on occurrence, distribution [J]. Marine Pollution Bulletin, 2014, 86(1-2): 562-568
[39] Minor E C, Lin R, Burrows A, et al. An analysis of microlitter and microplastics from Lake Superior beach sand and surface-water [J]. Science of the Total Environment, 2020, 744: 140824
[40] Kerrigan J F, Sandberg K D, Engstrom D R, et al. Sedimentary record of antibiotic accumulation in Minnesota Lakes [J]. Science of the Total Environment, 2018, 621: 970-979
[41] Zhu X P, Ran W, Teng J, et al. Microplastic pollution in nearshore sediment from the Bohai Sea coastline [J]. Bulletin of Environmental Contamination and Toxicology, 2021, 107(4): 665-670
[42] Lu J, Zhang Y X, Wu J, et al. Occurrence and spatial distribution of antibiotic resistance genes in the Bohai Sea and Yellow Sea areas, China [J]. Environmental Pollution, 2019, 252(Pt A): 450-460
[43] Li J, Zhang K N, Zhang H. Adsorption of antibiotics on microplastics [J]. Environmental Pollution, 2018, 237: 460-467
[44] Yang Y Y, Liu G H, Song W J, et al. Plastics in the marine environment are reservoirs for antibiotic and metal resistance genes [J]. Environment International, 2019, 123: 79-86
[45] 范秀磊, 甘容, 谢雅, 等. 老化前后聚乳酸和聚乙烯微塑料对抗生素的吸附解吸行为[J]. 环境科学研究, 2021, 34(7): 1747-1756
Fan X L, Gan R, Xie Y, et al. Adsorption and desorption behavior of antibiotics on polylactic acid and polyethylene microplastics before and after aging [J]. Research of Environmental Sciences, 2021, 34(7): 1747-1756 (in Chinese)
[46] Antony A, Fudianto R, Cox S, et al. Assessing the oxidative degradation of polyamide reverse osmosis membrane—Accelerated ageing with hypochlorite exposure [J]. Journal of Membrane Science, 2010, 347(1-2): 159-164
[47] Yu F, Li Y, Huang G Q, et al. Adsorption behavior of the antibiotic levofloxacin on microplastics in the presence of different heavy metals in an aqueous solution [J]. Chemosphere, 2020, 260: 127650
[48] Liu G Z, Zhu Z L, Yang Y X, et al. Sorption behavior and mechanism of hydrophilic organic chemicals to virgin and aged microplastics in freshwater and seawater [J]. Environmental Pollution, 2019, 246: 26-33
[49] Yang C F, Guan J N, Yang Y D, et al. Interface behavior changes of weathered polystyrene with ciprofloxacin in seawater environment [J]. Environmental Research, 2022, 212(Pt A): 113132
[50] Guo X, Chen C, Wang J L. Sorption of sulfamethoxazole onto six types of microplastics [J]. Chemosphere, 2019, 228: 300-308
[51] Ziccardi L M, Edgington A, Hentz K, et al. Microplastics as vectors for bioaccumulation of hydrophobic organic chemicals in the marine environment: A state-of-the-science review [J]. Environmental Toxicology and Chemistry, 2016, 35(7): 1667-1676
[52] Puckowski A, CwięK W, Mioduszewska K, et al. Sorption of pharmaceuticals on the surface of microplastics [J]. Chemosphere, 2021, 263: 127976
[53] Guo X T, Pang J W, Chen S Y, et al. Sorption properties oftylosin on four different microplastics [J]. Chemosphere, 2018, 209: 240-245
[54] Xu B L, Liu F, Brookes P C, et al. The sorption kinetics and isotherms of sulfamethoxazole with polyethylene microplastics [J]. Marine Pollution Bulletin, 2018, 131(Pt A): 191-196
[55] Ding L, Mao R F, Ma S R, et al. High temperature depended on the ageing mechanism of microplastics under different environmental conditions and its effect on the distribution of organic pollutants [J]. Water Research, 2020, 174: 115634
[56] Liu P, Qian L, Wang H Y, et al. New insights into the aging behavior of microplastics accelerated by advanced oxidation processes [J]. Environmental Science &Technology, 2019, 53(7): 3579-3588
[57] Razanajatovo R M, Ding J N, Zhang S S, et al. Sorption and desorption of selected pharmaceuticals by polyethylene microplastics [J]. Marine Pollution Bulletin, 2018, 136: 516-523
[58] Xu B L, Liu F, Brookes P C, et al. Microplastics play a minor role in tetracycline sorption in the presence of dissolved organic matter [J]. Environmental Pollution, 2018, 240: 87-94
[59] Fu J X, Li Y N, Peng L, et al. Distinct chemical adsorption behaviors of sulfanilamide as a model antibiotic onto weathered microplastics in complex systems [J]. Colloids and Surfaces A: Physicochemical and Engineering Aspects, 2022, 648: 129337
[60] Sun M, Yang Y K, Huang M L, et al. Adsorption behaviors and mechanisms of antibiotic norfloxacin on degradable and nondegradable microplastics [J]. The Science of the Total Environment, 2022, 807(Pt 3): 151042
[61] Guo X, Liu Y, Wang J L. Sorption of sulfamethazine onto different types of microplastics: A combined experimental and molecular dynamics simulation study [J]. Marine Pollution Bulletin, 2019, 145: 547-554
[62] Velzeboer I, Kwadijk C J A F, Koelmans A A. Strong sorption of PCBs to nanoplastics, microplastics, carbon nanotubes, and fullerenes [J]. Environmental Science &Technology, 2014, 48(9): 4869-4876
[63] Chen Y J, Li J N, Wang F H, et al. Adsorption of tetracyclines onto polyethylene microplastics: A combined study of experiment and molecular dynamics simulation [J]. Chemosphere, 2021, 265: 129133
[64] Zhang H B, Wang J Q, Zhou B Y, et al. Enhanced adsorption of oxytetracycline to weathered microplastic polystyrene: Kinetics, isotherms and influencing factors [J]. Environmental Pollution, 2018, 243(Pt B): 1550-1557
[65] Atugoda T, Wijesekara H, Werellagama D R I B, et al. Adsorptive interaction of antibiotic ciprofloxacin on polyethylene microplastics: Implications for vector transport in water [J]. Environmental Technology &Innovation, 2020, 19: 100971
[66] Zhou Z Q, Sun Y R, Wang Y Y, et al. Adsorption behavior of Cu(Ⅱ) and Cr(Ⅵ) on aged microplastics in antibiotics-heavy metals coexisting system [J]. Chemosphere, 2022, 291(Pt 1): 132794
[67] Wu P F, Cai Z W, Jin H B, et al. Adsorption mechanisms of five bisphenol analogues on PVC microplastics [J]. The Science of the Total Environment, 2019, 650(Pt 1): 671-678
[68] Yamate T, Kumazawa K, Suzuki H, et al. CH/π interactions for macroscopic interfacial adhesion design [J]. ACS Macro Letters, 2016, 5(7): 858-861
[69] Kong F X, Xu X, Xue Y G, et al. Investigation of the adsorption of sulfamethoxazole by degradable microplastics artificially aged by chemical oxidation [J]. Archives of Environmental Contamination and Toxicology, 2021, 81(1): 155-165
[70] Atugoda T, Vithanage M, Wijesekara H, et al. Interactions between microplastics, pharmaceuticals and personal care products: Implications for vector transport [J]. Environment International, 2021, 149: 106367
[71] Xiong Y C, Zhao J H, Li L Q, et al. Interfacial interaction between micro/nanoplastics and typical PPCPs and nanoplastics removal via electrosorption from an aqueous solution [J]. Water Research, 2020, 184: 116100
[72] Fan X L, Zou Y F, Geng N, et al. Investigation on the adsorption and desorption behaviors of antibiotics by degradable MPs with or without UV ageing process [J]. Journal of Hazardous Materials, 2021, 401: 123363
[73] González-Pleiter M, Pedrouzo-Rodríguez A, Verdú I, et al. Microplastics as vectors of the antibiotics azithromycin and clarithromycin: Effects towards freshwater microalgae [J]. Chemosphere, 2021, 268: 128824
[74] Godoy V, Martín-Lara M A, Calero M, et al. The relevance of interaction of chemicals/pollutants and microplastic samples as route for transporting contaminants [J]. Process Safety and Environmental Protection, 2020, 138: 312-323
[75] Yao J J, Wen J Y, Li H P, et al. Surface functional groups determine adsorption of pharmaceuticals and personal care products on polypropylene microplastics [J]. Journal of Hazardous Materials, 2022, 423(Pt B): 127131
[76] Hüffer T, Hofmann T. Sorption of non-polar organic compounds by micro-sized plastic particles in aqueous solution [J]. Environmental Pollution, 2016, 214: 194-201
[77] 郭梦函. 抗生素在微塑料上的吸附行为及其相关毒性研究[D]. 西安: 西安理工大学, 2020: 19-21
Guo M H. Studies on the adsorption behavior of antibiotics on microplastics and its related toxicity [D]. Xi’an: Xi’an University of Technology, 2020: 19-21 (in Chinese)
[78] Syranidou E, Kalogerakis N. Interactions of microplastics, antibiotics and antibiotic resistant genes within WWTPs [J]. Science of the Total Environment, 2022, 804: 150141
[79] Santos L H M L M, Rodríguez-Mozaz S, Barceló D. Microplastics as vectors of pharmaceuticals in aquatic organisms - An overview of their environmental implications [J]. Case Studies in Chemical and Environmental Engineering, 2021, 3: 100079
[80] 孔凡星, 许霞, 薛银刚, 等. 微塑料老化对四环素吸附行为的影响[J]. 环境科学研究, 2021, 34(9): 2182-2190
Kong F X, Xu X, Xue Y G, et al. Effect of aging on adsorption of tetracycline by microplastics [J]. Research of Environmental Sciences, 2021, 34(9): 2182-2190 (in Chinese)
[81] Li X N, Chen S, Fan X F, et al. Adsorption of ciprofloxacin, bisphenol and 2-chlorophenol on electrospun carbon nanofibers: In comparison with powder activated carbon [J]. Journal of Colloid and Interface Science, 2015, 447: 120-127
[82] Brennecke D, Duarte B, Paiva F, et al. Microplastics as vector for heavy metal contamination from the marine environment [J]. Estuarine, Coastal and Shelf Science, 2016, 178: 189-195
[83] Wang F, Shih K M, Li X Y. The partition behavior of perfluorooctanesulfonate (PFOS) and perfluorooctanesulfonamide (FOSA) on microplastics [J]. Chemosphere, 2015, 119: 841-847
[84] Elgarahy A M, Akhdhar A, Elwakeel K Z. Microplastics prevalence, interactions, and remediation in the aquatic environment: A critical review [J]. Journal of Environmental Chemical Engineering, 2021, 9(5): 106224
[85] Fu L N, Li J, Wang G Y, et al. Adsorption behavior of organic pollutants on microplastics [J]. Ecotoxicology and Environmental Safety, 2021, 217: 112207
[86] Wang J, Liu X H, Liu G N, et al. Size effect of polystyrene microplastics on sorption of phenanthrene and nitrobenzene [J]. Ecotoxicology and Environmental Safety, 2019, 173: 331-338
[87] Arienzo M, Ferrara L, Trifuoggi M. The dual role of microplastics in marine environment: Sink and vectors of pollutants [J]. Journal of Marine Science and Engineering, 2021, 9(6): 642
[88] Sun P P, Liu X M, Zhang M H, et al. Sorption and leaching behaviors between aged MPs and BPA in water: The role of BPA binding modes within plastic matrix [J]. Water Research, 2021, 195: 116956
[89] Gong W W, Jiang M Y, Han P, et al. Comparative analysis on the sorption kinetics and isotherms of fipronil on nondegradable and biodegradable microplastics [J]. Environmental Pollution, 2019, 254(Pt A): 112927
[90] Elizalde-Velázquez A, Subbiah S, Anderson T A, et al. Sorption of three common nonsteroidal anti-inflammatory drugs (NSAIDs) to microplastics [J]. The Science of the Total Environment, 2020, 715: 136974
[91] Yu X X, Du H H, Huang Y H, et al. Selective adsorption of antibiotics on aged microplastics originating from mariculture benefits the colonization of opportunistic pathogenic bacteria [J]. Environmental Pollution, 2022, 313: 120157
[92] Ajouyed O, Hurel C, Ammari M, et al. Sorption of Cr(Ⅵ) onto natural iron and aluminum (oxy)hydroxides: Effects of pH, ionic strength and initial concentration [J]. Journal of Hazardous Materials, 2010, 174(1-3): 616-622
[93] Ma J, Zhao J H, Zhu Z L, et al. Effect of microplastic size on the adsorption behavior and mechanism of triclosan on polyvinyl chloride [J]. Environmental Pollution, 2019, 254(Pt B): 113104
[94] Aristilde L, Marichal C, Miéhé-Brendlé J, et al. Interactions of oxytetracycline with a smectite clay: A spectroscopic study with molecular simulations [J]. Environmental Science &Technology, 2010, 44(20): 7839-7845
[95] Guo X, Wang J L. Sorption of antibiotics onto aged microplastics in freshwater and seawater [J]. Marine Pollution Bulletin, 2019, 149: 110511
[96] Turku I, Sainio T, Paatero E. Thermodynamics of tetracycline adsorption on silica [J]. Environmental Chemistry Letters, 2007, 5(4): 225-228
[97] Xue X D, Hong S C, Cheng R T, et al. Adsorption characteristics of antibiotics on microplastics: The effect of surface contamination with an anionic surfactant [J]. Chemosphere, 2022, 307(Pt 4): 136195
[98] 于明睿, 万涛, 熊静, 等. 磁性聚合物吸附剂的制备及其重金属离子吸附性能[J]. 功能材料与器件学报, 2022, 28(2): 166-171
Yu M R, Wan T, Xiong J, et al. Synthesis and adsorption properties of magnetic polymer adsorbent for heavy metal ions [J]. Journal of Functional Materials and Devices, 2022, 28(2): 166-171 (in Chinese)
[99] 李若男, 周丽莎, 陈舜胜, 等. 纤维素纳米纤维及其改性产物吸附重金属的研究进展[J]. 化工进展, 2022, 41(1): 310-319
Li R N, Zhou L S, Chen S S, et al. Research progress on adsorption of heavy metals by cellulose nanofibers and their modified products [J]. Chemical Industry and Engineering Progress, 2022, 41(1): 310-319 (in Chinese)
[100] 阮玉婷, 姚平, 陈国强, 等. 氧化石墨烯基材料吸附重金属离子的研究进展[J]. 印染助剂, 2022, 39(2): 9-17
Ruan Y T, Yao P, Chen G Q, et al. Research progress on adsorption of heavy metal ions by graphene oxide-based materials [J]. Textile Auxiliaries, 2022, 39(2): 9-17 (in Chinese)
[101] Wang F T, Pan Y F, Cai P X, et al. Single and binary adsorption of heavy metal ions from aqueous solutions using sugarcane cellulose-based adsorbent [J]. Bioresource Technology, 2017, 241: 482-490
[102] 王一飞. 微塑料对氟喹诺酮类抗生素的吸附作用[D]. 金华: 浙江师范大学, 2021: 14-15
Wang Y F. Adsorption of fluoroquinolones by microplastics [D]. Jinhua: Zhejiang Normal University, 2021: 14-15 (in Chinese)
[103] Barboza L G A, DickVethaak A, Lavorante B R B O, et al. Marine microplastic debris: An emerging issue for food security, food safety and human health [J]. Marine Pollution Bulletin, 2018, 133: 336-348
[104] Barboza L G A, Vieira L R, Guilhermino L. Single and combined effects of microplastics and mercury on juveniles of the European seabass (Dicentrarchus labrax): Changes in behavioural responses and reduction of swimming velocity and resistance time [J]. Environmental Pollution, 2018, 236: 1014-1019
[105] Wright S L, Kelly F J. Plastic and human health: A micro issue? [J]. Environmental Science &Technology, 2017, 51(12): 6634-6647
[106] 屈沙沙, 朱会卷, 刘锋平, 等. 微塑料吸附行为及对生物影响的研究进展[J]. 环境卫生学杂志, 2017, 7(1): 75-78
Qu S S, Zhu H J, Liu F P, et al. Adsorption behavior and effect on biont of microplastic [J]. Journal of Environmental Hygiene, 2017, 7(1): 75-78 (in Chinese)
[107] Oliveira R, McDonough S, Ladewig J C L, et al. Effects of oxytetracycline and amoxicillin on development and biomarkers activities of zebrafish (Danio rerio) [J]. Environmental Toxicology and Pharmacology, 2013, 36(3): 903-912
[108] Almeida A R,Tacão M, Machado A L, et al. Long-term effects of oxytetracycline exposure in zebrafish: A multi-level perspective [J]. Chemosphere, 2019, 222: 333-344
[109] Zhou L, Limbu S M, Qiao F, et al. Influence of long-term feeding antibiotics on the gut health of zebrafish [J]. Zebrafish, 2018, 15(4): 340-348
[110] Zhou L, Limbu S M, Shen M L, et al. Environmental concentrations of antibiotics impair zebrafish gut health [J]. Environmental Pollution, 2018, 235: 245-254
[111] Petersen B D, Pereira T C B, Altenhofen S, et al. Antibiotic drugs alter zebrafish behavior [J]. Comparative Biochemistry and Physiology Toxicology &Pharmacology, 2021, 242: 108936
[112] Zhang Q, Cheng J P, Xin Q. Effects of tetracycline on developmental toxicity and molecular responses in zebrafish (Danio rerio) embryos [J]. Ecotoxicology, 2015, 24(4): 707-719
[113] Keerthisinghe T P, Wang F, Wang M J, et al. Long-term exposure to TET increases body weight of juvenile zebrafish as indicated in host metabolism and gut microbiome [J]. Environment International, 2020, 139: 105705
[114] Qiu W H, Liu X J, Yang F, et al. Single and joint toxic effects of four antibiotics on some metabolic pathways of zebrafish (Danio rerio) larvae [J]. The Science of the Total Environment, 2020, 716: 137062
[115] Gonçalves C L, Vasconcelos F F P, Wessler L B, et al. Exposure to a high dose of amoxicillin causes behavioral changes and oxidative stress in young zebrafish [J]. Metabolic Brain Disease, 2020, 35(8): 1407-1416
[116] Xi J L, Liu J, He S J, et al. Effects of norfloxacin exposure on neurodevelopment of zebrafish (Danio rerio) embryos [J]. Neurotoxicology, 2019, 72: 85-94
[117] Shen R, Yu Y C, Lan R, et al. The cardiovascular toxicity induced by high doses of gatifloxacin and ciprofloxacin in zebrafish [J]. Environmental Pollution, 2019, 254(Pt B): 112861
[118] Yan Z H, Lu G H, Ye Q X, et al. Long-term effects of antibiotics, norfloxacin, and sulfamethoxazole, in a partial life-cycle study with zebrafish (Danio rerio): Effects on growth, development, and reproduction [J]. Environmental Science and Pollution Research International, 2016, 23(18): 18222-18228
[119] Minski V T, Garbinato C, Thiel N, et al. Erythromycin in the aquatic environment: Deleterious effects on the initial development of zebrafish [J]. Journal of Toxicology and Environmental Health Part A, 2021, 84(2): 56-66
[120] He J H, Guo S Y, Zhu F, et al. A zebrafish phenotypic assay for assessing drug-induced hepatotoxicity [J]. Journal of Pharmacological and Toxicological Methods, 2013, 67(1): 25-32
[121] Zhang M Q, Chen B, Zhang J P, et al. Liver toxicity of macrolide antibiotics in zebrafish [J]. Toxicology, 2020, 441: 152501
[122] Yan Z Y, Huang X Y, Xie Y, et al. Macrolides induce severe cardiotoxicity and developmental toxicity in zebrafish embryos [J]. The Science of the Total Environment, 2019, 649: 1414-1421
[123] Han E, Oh K H, Park S, et al. Analysis of behavioral changes in zebrafish (Danio rerio) larvae caused by aminoglycoside-induced damage to the lateral line and muscles [J]. Neurotoxicology, 2020, 78: 134-142
[124] Ton C, Parng C. The use of zebrafish for assessing ototoxic and otoprotective agents [J]. Hearing Research, 2005, 208(1-2): 79-88
[125] Shao W H, Zhong D, Jiang H W, et al. A new aminoglycoside etimicin shows low nephrotoxicity and ototoxicity in zebrafish embryos [J]. Journal of Applied Toxicology, 2021, 41(7): 1063-1075
[126] Hentschel D M, Park K M, Cilenti L, et al. Acute renal failure in zebrafish: A novel system to study a complex disease [J]. American Journal of Physiology Renal Physiology, 2005, 288(5): F923-F929
[127] Liu J Y, Wei T Z, Wu X, et al. Early exposure to environmental levels of sulfamethoxazole triggers immune and inflammatory response of healthy zebrafish larvae [J]. The Science of the Total Environment, 2020, 703: 134724
[128] Yan Z Y, Yang Q L, Jiang W L, et al. Integrated toxic evaluation of sulfamethazine on zebrafish: Including two lifespan stages (embryo-larval and adult) and three exposure periods (exposure, post-exposure and re-exposure) [J]. Chemosphere, 2018, 195: 784-792
[129] Yang L Y, Higginbotham J N, Liu L P, et al. Production of a functional factor, p40, by Lactobacillus rhamnosus GG is promoted by intestinal epithelial cell-secreted extracellular vesicles [J]. Infection and Immunity, 2019, 87(7): e00113-e00119
[130] Jin Y X, Wu S S, Zeng Z Y, et al. Effects of environmental pollutants on gut microbiota [J]. Environmental Pollution, 2017, 222: 1-9
[131] Liu S L, Yan L, Zhang Y L, et al. Polystyrene nanoplastics exacerbated the ecotoxicological and potential carcinogenic effects of tetracycline in juvenile grass carp (Ctenopharyngodon idella) [J]. The Science of the Total Environment, 2022, 803: 150027
[132] Zhang P, Lu G H, Sun Y, et al. Aged microplastics change the toxicological mechanism of roxithromycin on Carassius auratus: Size-dependent interaction and potential long-term effects [J]. Environment International, 2022, 169: 107540
[133] Zhang P, Lu G H, Sun Y, et al. Metagenomic analysis explores the interaction of aged microplastics and roxithromycin on gut microbiota and antibiotic resistance genes of Carassius auratus [J]. Journal of Hazardous Materials, 2022, 425: 127773
[134] Lu J, Zhang Y X, Wu J, et al. Effects of microplastics on distribution of antibiotic resistance genes in recirculating aquaculture system [J]. Ecotoxicology and Environmental Safety, 2019, 184: 109631
[135] Bakir A, Rowland S J, Thompson R C. Enhanced desorption of persistent organic pollutants from microplastics under simulated physiological conditions [J]. Environmental Pollution, 2014, 185: 16-23
[136] Zhang Y T, Chen H X, He S Q, et al. Subchronic toxicity of dietary sulfamethazine and nanoplastics in marine medaka (Oryzias melastigma): Insights from the gut microbiota and intestinal oxidative status [J]. Ecotoxicology and Environmental Safety, 2021, 226: 112820
[137] Liao X, Zhao P Q, Hou L Y, et al. Network analysis reveals significant joint effects of microplastics and tetracycline on the gut than the gill microbiome of marine medaka [J]. Journal of Hazardous Materials, 2023, 442: 129996
[138] Zhang S S, Ding J N, Razanajatovo R M, et al. Interactive effects of polystyrene microplastics and roxithromycin on bioaccumulation and biochemical status in the freshwater fish red tilapia (Oreochromis niloticus) [J]. The Science of the Total Environment, 2019, 648: 1431-1439
[139] Broom D M. Fish brains and behaviour indicate capacity for feeling pain [J]. Animal Sentience, 2016, 1(3): 4
[140] Dvir H, Silman I, Harel M, et al. Acetylcholinesterase: From 3D structure to function [J].Chemico-Biological Interactions, 2010, 187(1-3): 10-22
[141] Huang Y J, Ding J N, Zhang G S, et al. Interactive effects of microplastics and selected pharmaceuticals on red tilapia: Role of microplastic aging [J]. The Science of the Total Environment, 2021, 752: 142256
[142] Lu J R, Wu J, Gong L L, et al. Combined toxicity of polystyrene microplastics and sulfamethoxazole on zebrafish embryos [J]. Environmental Science and Pollution Research International, 2022, 29(13): 19273-19282
[143] Segner H. Reproductive and Developmental Toxicity in Fishes [M]//Reproductive and Developmental Toxicology. Amsterdam: Elsevier, 2011: 1145-1166
[144] Qiu W H, Fang M J, Magnuson J T, et al. Maternal exposure to environmental antibiotic mixture during gravid period predicts gastrointestinal effects in zebrafish offspring [J]. Journal of Hazardous Materials, 2020, 399: 123009
[145] Cormier B,Cachot J, Blanc M, et al. Environmental microplastics disrupt swimming activity in acute exposure in Danio rerio larvae and reduce growth and reproduction success in chronic exposure in D. rerio and Oryzias melastigma [J]. Environmental Pollution, 2022, 308: 119721
[146] He S Q, Li D, Wang F P, et al. Parental exposure to sulfamethazine and nanoplastics alters the gut microbial communities in the offspring of marine medaka (Oryzias melastigma) [J]. Journal of Hazardous Materials, 2022, 423(Pt A): 127003
[147] Park S H, Kim K. Microplastics induced developmental toxicity with microcirculation dysfunction in zebrafish embryos [J]. Chemosphere, 2022, 286(Pt 3): 131868
[148] Yang Q L, Gao Y, Ke J, et al. Antibiotics: An overview on the environmental occurrence, toxicity, degradation, and removal methods [J]. Bioengineered, 2021, 12(1): 7376-7416
[149] Fonte E, Ferreira P, Guilhermino L. Temperature rise and microplastics interact with the toxicity of the antibiotic cefalexin to juveniles of the common goby (Pomatoschistus microps): Post-exposure predatory behaviour, acetylcholinesterase activity and lipid peroxidation [J]. Aquatic Toxicology, 2016, 180: 173-185