2023, Vol. 44
2. 重庆大学煤矿灾害动力学与控制国家重点实验室, 重庆 400044
2. State Key Laboratory of Coal Mine Disaster Dynamics and Control, Chongqing University, Chongqing 400044, China
季铵盐(quaternary ammonium compounds, QACs)是一类由一个带正电的中心N+和其连接的有机基团以及带负电的卤素原子构成的阳离子杀菌剂, 主要通过烷基链改变细胞膜的磷脂双分子层, 破坏细胞膜, 使细胞膜内的物质流失导致细菌死亡[1, 2].根据其化学结构式, QACs主要分为三大类(如图 1)：①直链烷基铵类、②咪唑类和③吡啶类[3].一般地, 氯化QACs杀菌效果强于其他QACs[4], 含苯环QACs强于不含苯环类[5].其中, 直链烷基铵类的烷基三甲基铵盐(alkyltrimethylammonium compounds, ATMACs)、二烷基二甲基铵盐(dialkyldimethylammonium compounds, DADMACs)和苄基烷基二甲基铵盐(benzylammonium compounds, BACs)是最常用的QACs类杀菌剂.
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图 1 典型QACs结构式 Fig. 1 Structural formulas of typical QACs |
由于其良好的杀菌性能和双亲性, QACs被广泛应用在工业、农业和医护行业[6~8].美国环保署在2015年公布的报告中指出ATMACs、DADMACs和BACs在美国5年内总产量约4 536~22 680 t(1 000万~5 000万磅)[9].近年来, 流感和新冠肺炎大流行, 更高的消毒要求导致了更多杀菌剂生产和使用[10, 11].有数据表明, 用于对抗新冠病毒的杀菌剂中一半以上是以QACs作为活性成分[12].QACs在使用过程中会有意或无意释放到环境中, 据测算, 约75%的QACs被排放到污水处理系统, 其余则直接释放到环境中[13].由于传统污水处理技术的局限性, QACs在污水处理厂不能被完全去除, 如澳大利亚某市政污水处理厂进水中ATMAC(C16)浓度范围为7.7~27 μg·L-1, 出水浓度约为1.1 μg·L-1[14].同时, QACs的正电性和疏水性导致其易吸附在活性污泥中, 最终随污水处理厂出水或活性污泥农用释放到环境中.如墨西哥一处利用污水处理厂出水灌溉的农田土壤中BAC(C12)浓度为81 μg·kg-1[15], 显著高于其他未灌溉区土壤.目前QACs在世界范围内的水体、沉积物和土壤中均有检出, 在水中浓度范围为几~几十μg·L-1[16~18], 在沉积物和土壤中浓度范围为几十~几千μg·kg-1[15, 19~21]. QACs在环境中的浓度低于其杀菌浓度(10~50 mg·L-1)[22], 且极易吸附在微生物体、土壤和沉积物等介质表面, 生物可利用性降低, 导致QACs在环境中的赋存时间延长.长期低剂量的QACs暴露易诱导微生物对其产生抗性, 导致QACs杀菌失效.据统计, 已有16起流行疾病的暴发是由于微生物出现抗性导致QACs杀菌剂失效引起的[23].
由于大量生产和使用, QACs广泛存在于多种环境介质中, 对生态系统和人体健康构成潜在威胁, 了解其环境行为及效应是实现有效管控的基础.本文结合国内外相关文献, 综述了QACs在环境中的迁移转化行为、毒性效应、微生物出现QACs抗性及诱导抗生素抗性的主要机制, 以期为充分了解QACs的环境风险提供重要参考.
1 QACs在环境中的主要迁移转化过程QACs在环境中的迁移转化方式主要包括吸附和降解, 其在活性污泥、土壤和沉积物等固相介质中的吸附分为可逆吸附和不可逆吸附, 可逆吸附受降解过程影响.QACs在环境中的降解主要包括光解和生物降解.目前仅有一篇关于QACs在水体中光解的研究, 发现QACs能够吸收的太阳光谱有限, 自然光解缓慢, 半衰期为12~94 d, 相比非生物降解, 生物降解是QACs在环境中的主要消除方式[24].在厌氧条件下, QACs很难被生物降解利用, 而在好氧条件下可以作为碳源被生物降解利用, 半衰期为数小时到十几天[25, 26].
1.1 QACs的主要吸附特征QACs具有带正电的头部(N+)和长的疏水性尾部(碳氢链), 进入环境后容易通过静电作用和疏水作用吸附在微生物、活性污泥、土壤和沉积物等带负电的介质上[27], 目前的研究主要聚焦于QACs在活性污泥和土壤中的吸附机制.
活性污泥对QACs的吸附主要是通过有机质的疏水作用实现的, 吸附强弱与QACs结构有关, 一般QACs烷基链越长越易被吸附[28~31].有研究发现活性污泥对4种QACs的吸附能力依次为：ATMAC(C16)>BAC(C16)>BAC(C12)>ATMAC(C12), 吸附亲和力与烷基链长呈正相关; QACs烷基链较短时, 苄基能增加吸附亲和力[29].此外, pH也是重要的影响因素, 随着pH值的升高, 活性污泥对QACs的吸附量降低, 主要是由于高pH会导致活性污泥中的有机质被大量溶解, 产生大量带负电的酸性基团, 这些酸性基团与QACs发生静电作用, 从而抑制了活性污泥对QACs的吸附[29, 32].
土壤是一个由微生物、黏土矿物、腐殖质及其他固体和可溶性物质组成的复杂有机-无机复合体.土壤中的有机质、黏土矿物、氧化物和微生物细胞壁均可以作为QACs的吸附剂.Sarkar等[33]研究了QACs在3种土壤中的吸附行为, 发现土壤矿质黏土成分是影响QACs在土壤中吸附强度的主要因素, 蒙脱石对QACs的吸附强度大于高岭土.进一步, 土壤矿质胶体性质也会对QACs吸附产生影响, 土壤矿质胶体吸附的阳离子可与带正电的QACs进行离子交换, 快速达到吸附平衡,
与QACs浓度无关, 但当QACs浓度高于土壤阳离子交换容量(cation exchange capacity, CEC)的80%时, 由于空间位阻效应, 土壤吸附QACs速率降低[34].Ilari等[35]考察了不同离子浓度(0.002、0.02和0.1 mol·L-1 NaCl)对蒙脱土吸附BAC(C12)的影响特征, 发现最大吸附容量与离子浓度呈正相关.Ndabambi等[36]发现土壤对不同链长的BAC(C8~C14)的吸附能力与土壤CEC含量和BAC烷基碳链长度呈正相关, 表明土壤对BAC的吸附过程主要受离子交换和疏水作用影响.土壤的铁铝等氧化物会改变黏土矿物的表面积和净电荷零点, 影响土壤团聚体的稳定, 改变土壤对QACs的吸附, 氧化铁Zata电位的改变也会加强或抑制吸附, 关于土壤中氧化物对QACs吸附机制的研究有待深入.土壤微生物是土壤最活跃的组分之一且带负电荷, 也会影响土壤对QACs的吸附性, 同时, 吸附态的QACs也会影响其在土壤中的生物有效性[33].
综上, QACs在环境中的吸附过程主要受QACs自身理化性质和环境介质理化性质影响, 吸附机制主要涉及离子交换、静电作用及疏水作用等.
1.2 QACs的生物降解特征虽然环境介质对QACs有较强的吸附性, 但在目前报道的机制中, 生物降解才是其在环境中自然衰减的主要方式.有研究模拟了10种QACs在自然水环境中的生物转化, 发现其均能被生物降解, 半衰期为0.5~1.6 d[37].本课题组研究了BAC(C12)在两种不同类型土壤(碱性耕地土壤和酸性森林土壤)中的好氧生物降解, 降解半衰期分别为4.66 d和17.33 d, 土壤有机质含量和微生物群落结构对其降解速率存在显著影响[38].
QACs的好氧生物降解主要由于Xanthomonas、Aeromonas和Pseudomonas等功能菌的生长利用.Oh等[26]将BAC(C12和C14)作为好氧间歇式反应器的唯一碳源, 发现12 h内大部分的BAC(C12和C14)被降解, Pseudomonas丰度显著上升.目前, 研究者在好氧条件下分离出一些有效降解QACs的微生物, 如Pseudomonas fluorescens TN4、Pseudomonas putida A (ATCC 12633)和Aeromonas hydrophila sp. K等[39~41].
QACs的好氧生物降解反应主要由不同位置碳的羟基化来起始, 目前报道的QACs好氧生物降解主要有以下3条途径(如图 2)：①烷基末端C原子的ω-羟基化起始反应, 后经脱乙酸作用和末端羟基氧化形成羧酸-QACs, 后经脱羧反应及多次脱甲基反应降解为NH4+、CO2和H2O, 乙酸和羧酸基团则通过β-氧化被降解为CO2和H2O; ②烷基上与N原子相邻的C原子的α-羟基化起始反应, 后经脱烷基作用生成叔铵化合物和长链羧酸, 叔铵化合物经过多次脱甲基后产生NH4+、CO2和H2O, 长链羧酸则通过β-氧化被降解为CO2和H2O; ③甲基C原子的α-羟基化起始反应, 然后脱甲酸, 之后按途径②被降解为CO2和H2O[13, 42, 43].3种途径的起始能量负荷是相同的, 但途径②产生的中间产物(叔铵化合物)链长更短, 比途径①和③的产物毒性小, 因此, 途径②是主要的生物转化机制[13].在降解过程中微生物的单加氧酶、胺氧化酶和环羟基化加氧酶发挥关键作用[26, 44, 45].上述途径均是纯菌株降解利用QACs的主要方式, 在多种微生物共存、多种电子受体共存的实际环境介质中, QACs的主要降解途径是否会发生变化, 环境中的多种理化因子是否对其产生影响均有待进一步研究.
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图 2 QACs的好氧生物降解途径 Fig. 2 Aerobic biodegradation pathway of QACs |
作为一种非专一性杀菌剂, 环境中残留的QACs会对鱼类、藻类和微生物等多种生物产生毒性效应, 尤其是会诱发细菌出现抗性, 导致杀菌失效, 有引发传染病流行的潜在风险.本部分主要对目前关于QACs生物毒性的研究进行归纳总结, 便于研究者深入理解其可能带来的潜在生态风险.
2.1 急性毒性QACs对于土壤微生物的急性毒性浓度往往高达数百mg·kg-1.对土壤微生物硝化作用的48 h急性抑制试验发现, BAC半数效应浓度(EC50)值为221 mg·kg-146]; 对土壤中Bacillus cereus纯菌株生长抑制的EC50(48 h)为500 mg·kg-147].
表 1总结了目前报道的关于QACs对水生生物的急性毒性效应浓度.环境浓度下QACs不会对水生生物构成急性毒性, 其急性毒性效应浓度与其自身结构、受试生物类型及受试时间紧密相关.在所报道的受试生物中, Daphnia magna对QACs最敏感[48].如Kreuzinger等[17]对比了48 h内BAC(C12、C14和C16)对几种水生生物生长繁殖影响的EC50值, 发现BAC(C12、C14和C16)对Daphnia magna的EC50值为0.041 mg·L-1, 显著低于Brachionus calyciflorus(0.125 mg·L-1)和Tetrahymena thermophila(2.941 mg·L-1).一般地, QACs的急性毒性强度与其碳链长度成正比, 可能是因为长链QACs更容易被吸附到生物膜上, 破坏细胞膜结构.如96 h内, BAC(C12)、BAC(C14)和BAC(C16)对Chlorella vulgaris的生长抑制EC50值分别为0.203、0.174和0.161 mg·L-149].此外, 其毒性效应还受暴露时间影响, Li[50]将Dugesia japonica暴露在同一浓度BAC(C12)下, 发现半数致死浓度(LC50)与暴露时间呈反比.
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表 1 QACs对水生生物的急性毒性 Table 1 Acute toxicity of QACs to aquatic organisms |
为了进一步明确QACs对水生生物急性毒性机制, 本课题组利用差异分析蛋白组学探究了BAC(C12)对铜绿微囊藻的急性毒性效应及其机制, 发现96 h内BAC(C12)对Microcystis aeruginosa生长抑制的EC50为3.61 mg·L-151].在此剂量下, 藻的光合活性降低了36%, 但内源藻毒素和乳酸脱氢酶的胞外释放量显著增高, 藻细胞出现质壁分离、类囊体模糊及类囊体膜堆积等现象.基于差异分析蛋白组学分析, 表明BAC(C12)主要通过破坏光合系统, 导致电子传递受限, 构成氧化胁迫, 破坏细胞膜来抑制Microcystis aeruginosa的生长, 同时内源微囊藻毒素合成量上升, 并通过破损的细胞膜释放到环境中, 从而增加了BAC(C12)对水生生物的危害.
2.2 慢性毒性低于作用浓度的QACs不会直接杀死细菌, 但会抑制细菌活性.当进水BACs浓度达到5 mg·L-1时, 氨氧化细菌的活性受到抑制[58].Yang等[59]发现BAC(C12)会导致湖泊水体微生物的amoA和nifH基因的丰度下降, 表明BAC(C12)可能会影响微生物驱动的氮循环过程.除了影响细菌的氮循环功能外, QACs还会对微生物构成氧化胁迫, 有研究发现将好氧膜生物反应器暴露于十六烷基三甲基溴化铵(CTAB)后, 反应器中细菌细胞色素C氧化酶含量减少, 呼吸电子传递体系功能受阻, 胞内活性氧自由基含量上升[60, 61], 本课题组的研究也发现了BAC(C12)会对Microcystis aeruginosa构成氧化胁迫[51].
QACs对于Daphnia magna、Ceriodaphnia dubia和Rainbow trout等水生生物的慢性毒性效应浓度普遍在μg·L-1级.Lavorgna等[55]探究了BAC(C18)对Daphnia magna和Ceriodaphnia dubia生殖的影响, 发现21 d内, 0.06 μg·L-1的BAC(C18)抑制了Daphnia magna的生殖, 7 d内3.39 μg·L-1的BAC(C18)抑制了Ceriodaphnia dubia的生殖, 此两种受试生物常为鱼类等水生生物的食物, 在食物链中发挥关键作用, 低浓度QACs的长期慢性毒性效应需引起重视.
哺乳动物对于QACs的敏感性较低, 效应浓度普遍在数百mg·kg-162].有研究发现老鼠摄入120 mg·kg-1 QACs后, 其后代数量减少并且怀孕间隔时间增加[63].最近有学者在人体血液样本中检测到了不同链长的QACs, 并发现QACs会增加血液中炎症细胞因子水平, 抑制线粒体功能, 破坏胆固醇稳态, 效应与QACs浓度显著相关[64].
3 QACs诱导细菌出现QACs抗性的主要机制低于作用浓度的QACs会诱导细菌出现抗性, 如长期暴露在BACs环境中的Staphylococcus aureus会对BACs产生抗性, 最低抑菌浓度(minimum inhibitory concentration, MIC)从5 mg·L-1上升到10 mg·L-165].对单菌株的研究发现, 微生物主要通过修饰细胞膜结构、抑制膜孔蛋白对QACs的运输、药物外排泵蛋白高表达及捕获水平移动元件等机制获得QACs抗性.
细胞膜是QACs杀菌的主要靶位点, 细菌常通过改变细胞膜的结构和组成, 抑制QACs的进入, 降低对QACs的敏感性.例如, Pseudomonas aeruginosa BACs抗性株细胞膜中的磷脂和脂肪酸的含量显著高于敏感株[66], 在BACs诱导下, Bacillus cereus细胞膜中短链脂肪酸含量显著上升[67]; Pseudomonas aeruginosa通过增加亚精胺(一种聚阳离子)的合成来稳定膜电荷, 以减少QACs的进入[66].细菌还可以调控细胞膜运输蛋白的含量来降低对QACs的敏感性.有研究发现膜孔蛋白基因的下调能够降低Pseudomonas[68, 69]和Escherichia coli[70]对BACs的敏感性.
形成生物膜是细菌保护自身免受环境压力的常见抗性机制, 可以提高对QACs的抗性.例如, Staphylococcus epidermidis CIP53124形成生物膜后降低了对BACs的敏感性[71].从乳制品行业中分离出的BACs抗性Escherichia coli形成生物膜的能力强于敏感株, 敏感株暴露于BACs一段时间后其形成生物膜的能力也显著提升[72].此外, 多菌种生物膜对QACs的抗性强于单菌种生物膜[73, 74]. Giaouris等[74]发现多菌种生物膜的形成能够显著降低Pseudomonas putida对BACs的敏感性.
外排泵蛋白的过度表达也可以使细菌产生QACs抗性.外排泵是包含跨膜区的膜蛋白, 其形成的通道能够主动将物质从细胞质或者细胞膜中去除[75~77].有研究表明QACs抗性往往与Qac外排泵有关, Qac外排泵蛋白家族主要包括QacA、QacB、QacC、Qac EΔ1、QacG、QacH和QacJ, 其中QacA, QacB是主促进者(major facilitator, MF)家族, 而其余的则是小多药抗性(small multidrug resistance, SMR)家族[13, 78].有研究发现, Listeria monocytogenes通过QacH外排泵的高表达, 降低了对BACs敏感性, 而在外排泵抑制剂存在的条件下, 恢复了对BACs的敏感性[79, 80].此外, 细菌还可以借助抗生素类外排泵实现对QACs的外排.例如, 氟喹诺酮类抗生素外排泵CemABC、NorA和氨基糖苷类抗生素外排泵EmrE可协同将QACs外排[80, 81], 此类外排泵的过度表达, 会使细菌对QACs和相应的抗生素的耐受性提升2~8倍[82~85].
抗性基因的水平转移是细菌获得QACs抗性的另一重要途径, 许多QACs抗性基因通常位于质粒、整合子和转座子等可移动遗传元件(mobile genetic elements, MGEs)上.例如, 编码Qac外排泵的多种qac基因(qacA、qacB、qacC、qacEΔ1、qacG、qacH、qacJ)均位于质粒上[78]; 携带QACs抗性基因bcrABC的不相容质粒(IncP)广泛分布在QACs亚抑制浓度环境中, 且可以在所有的革兰氏阴性菌之间转移[86, 87]; 转座子Tn6188上也发现携带qacH和emrE基因[80].这些MGEs可以通过水平转移在不同细菌之间传播, 加剧了QACs抗性的扩散[88, 89].
在环境中, QACs的浓度远低于其MIC, 且QACs在环境中可以被好氧降解, 形成浓度梯度, 为细菌产生QACs抗性创造了良好的条件.Yang等[59]发现10 μg·L-1 QACs会诱导水体中微生物群落中qacA/B丰度急速上升.但实际环境中微生物应对QACs的主要抗性机制是什么、不同类型微生物的抗性机制是否存在差异、诱导环境微生物出现QACs抗性的最低浓度以及诱导QACs抗性基因水平转移的最低浓度是多少, 以上问题均有待揭示.
4 QACs对细菌抗生素抗性的协同选择机制由于作用机制和作用对象的相似性, QACs大量使用还可能诱导细菌出现抗生素抗性.细菌出现抗生素抗性将会导致更长的住院时间, 更高的死亡率, 以及更大的经济负担, 抗生素抗性菌的出现和传播已经成为一个日益严重的全球性问题[90~92].由于其广泛存在性、高传播性以及危害性, 抗生素抗性基因(antibiotic resistance genes, ARGs)已经被列为一类新型污染物.越来越多的证据表明, QACs会诱导细菌产生抗生素抗性.Han等[93]将从天然水体中分离出的Pseudomonas aeruginosa在含3种QACs的筛选培养基上培养30 d后, 发现其对四环素和环丙沙星的敏感性均显著降低.将湖泊水体微生物暴露在不同浓度的QACs下, 发现磺胺类抗性基因(sulⅠ)、四环素类抗性基因(tetA、tetM)和喹诺酮类抗性基因(qnrD)的丰度均有不同程度的上升[57].目前认为关于QACs对细菌抗生素抗性的选择机制主要包括协同抗性和交叉抗性.
协同抗性机制是指QACs抗性基因与ARGs位于同一MGEs上.多种QACs抗性基因和ARGs均位于Ⅰ类整合子(IntⅠ, 位于质粒或转座子上的DNA整合元件)上[94~96].Heuer等[97]对16种耕地土壤进行分析, 发现sulⅠ和qacEΔ1与IntⅠ成显著正相关.对Escherichia coli和Klebsiella pneumoniae进行基因组分析, 证实IntⅠ上携带β-内酰胺类抗生素抗性基因(blaIMP-11)、氨基糖苷类抗生素抗性基因(aacA1)、sulⅠ和qacEΔ1[98]. Han等[93]发现在48个水体样本中, 7种ARGs与IntⅠ和qacEΔ1显著正相关, 5种环境浓度下的QACs均显著促进了ARGs抗性质粒(RP4质粒)在大肠杆菌间的接合转移.有研究通过对2 666条公开可用的细菌染色体和1 926个质粒进行分析, 发现QACs抗性基因和ARGs之间存在正相关关系, qacEΔ1在质粒上被频繁检出[99].
交叉抗性是指细菌通过外排泵、形成生物膜和改变细胞膜组分等通用抗性机制获得QACs抗性的同时也获得抗生素抗性.如上文所述, 一些外排泵本身就是多药外排泵, 如NorA、CemABC和QacC外排泵也可同时将某些β-内酰胺类抗生素排出体外[100].生物膜是细菌保护自身免受环境压力的一种通用机制, McBain等[101]发现QACs能够促进细菌形成生物膜, 降低了细菌对QACs和抗生素的敏感性.同时, 细菌也可以通过改变细胞膜脂肪酸和磷脂的含量来降低细胞膜的通透性, 减少QACs和抗生素的摄入, 从而降低细菌的敏感性[102].
QACs导致微生物出现应激效应也会诱导ARGs的产生.Luo等[60]发现膜生物反应器中添加CTAB后, 多种ARGs的表达量升高, 其中四环素类抗性基因(tetO和tetW)表达量增加得最多.证实是由于CTAB损害了细胞膜和呼吸磷酸化, 干扰电子传递体系, 使得O2传递过程受阻, 细胞出现应激反应, 导致外排泵机制的ARGs过度表达[60, 103].
5 展望(1) 关于土壤和水等实际环境介质中QACs降解过程及微生物响应机制研究.实际环境中存在除分子氧外的其他多种电子受体, QACs的起始反应和降解机制有可能发生变化, 产生毒性更强的中间产物.进一步, 作为非专一性的杀菌剂, QACs及其降解产物会对微生物群落结构和相互作用关系产生怎样的影响？明确这些问题有助于明确QACs的环境行为和效应.
(2) 关于QACs对更多生物的长期慢性毒性效应研究, 尤其是在生态循环中发挥重要作用的生物.现有数据表明QACs在环境中的浓度不会对受试生物构成致死威胁, 但对Daphnia magna等水生物有显著慢性毒性效应, 且毒性效应浓度受生物种类影响显著.为了全面了解QACs的环境风险, 亟待分析其对更多生物的长期慢性毒性效应及其机制.进一步, QACs在环境中常与其他污染物共存, 其复合毒性效应研究也应给予充分关注.
(3) 关于环境微生物群落对QACs胁迫的主要抗性机制研究.单菌株的抗性机制研究发现, 由于细胞外膜结构的差异, 革兰氏阳性菌和革兰氏阴性菌对QACs抗性机制存在显著差异.在实际环境中, 多种微生物共存, 其应对QACs胁迫的主要抗性机制是什么, 不同类型微生物的应对策略是否存在差异以及它们之间的互作关系均不清晰.更重要的是, QACs是否会诱导环境微生物出现QACs抗性和促进抗性基因在不同种属间水平转移, 最低的作用浓度是多少？明确这些问题有助于制定有效的QACs管控措施.
6 结论(1) QACs在水体、土壤和沉积物中被广泛检出, 在水体中浓度范围为几~几十μg·L-1, 在沉积物和土壤中浓度范围为几十~几千μg·kg-1.吸附和好氧生物降解是其在环境中的主要迁移转化过程, 吸附机制主要涉及离子交换、静电作用和疏水作用.关于纯菌株对QACs降解的研究发现, 在好氧条件下, QACs可被Pseudomona、Aeromonas、Xanthomonas和Bacillus等种属细菌作为碳源或能源利用, 降解速度与微生物种属、共存离子显著相关.在分子氧和还原辅酶Ⅱ(NADPH)的参与下, 单加氧酶催化的羟基化反应为QACs生物降解的起始步骤, 差别仅在于发生羟基化的碳位点不同.
(2) 目前尚未发现环境浓度下QACs会对受试生物构成急性致死效应, 但环境浓度下QACs的长期暴露会抑制Daphnia magna的生长繁殖.QACs的毒性效应主要受自身结构、暴露时间和受试生物种类的显著影响.一般地, 烷基链越长毒性越强, 暴露时间越长效应浓度越低, 在已有关于水生生物研究中, Daphnia magna对QACs最敏感.
(3) 有机质和矿物质对QACs的可逆性吸附延长了其在环境中的赋存时间, 长期、低剂量QACs胁迫会诱导微生物出现QACs抗性.此外, QACs的生物转化导致其在环境中形成浓度梯度, 也利于微生物适应胁迫而出现抗性.除了降解利用, 微生物主要通过修饰细胞膜结构、抑制膜孔蛋白对QACs的输送、药物外排泵蛋白高表达、捕获MGEs获得QACs抗性基因以及形成生物膜等机制提高对QACs的抗性.此外, QACs还可以通过协同抗性和交叉抗性机制诱导微生物出现抗生素抗性, 加大环境中ARGs的传播风险.
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