1. 南京大学 环境学院 污染控制与资源化研究国家重点实验室，江苏 南京 210023;
2. 南京大学 泉州环保产业研究院，福建 泉州 362000
收稿日期：2021-06-03；接收日期：2021-08-18；网络出版时间：2021-09-01
基金项目：国家自然科学基金(Nos. 31861133003, 21806075)，泉州市科技计划项目(No. 2019C113) 资助
作者简介：季荣??南京大学教授、博士生导师，国家有机毒物污染控制与资源化工程技术研究中心副主任。德国康斯坦茨大学博士毕业，亚琛工业大学和马普研究所博士后。研究方向为污染物环境过程与效应。主持国家自然科学基金重点项目、面上项目及国际合作项目，科技部863项目(探索类)、重点研发计划课题，欧盟FP7和H2020等国际合作项目；发表论文200余篇，获授权国家发明专利10余项；现为中国化学会环境化学专业委员会秘书、中国环境学会环境化学专业委员会委员、中国毒理学会分析毒理专业委员会委员、中国土壤学会土壤生态专业委员会委员、中国土壤学会微塑料工作组副主任.
摘要：抗生素的广泛使用导致其在环境中普遍存在，所引发的抗性基因问题已对全球公共卫生构成重大威胁。土壤是环境中抗生素的重要汇，抗生素暴露会对土壤生物带来危害，甚至会间接对人体健康造成潜在风险，因此需采取有效手段修复抗生素污染的土壤。文中综述了抗生素对土壤植物表型生长指标、土壤动物生理特征及群落分布、微生物群落组成与功能的影响，以及抗生素抗性基因在土壤生物间的传播风险等；总结了利用耐受土壤植物、动物、微生物以及其互作关系修复抗生素污染土壤的潜力与前景，指出了已有土壤中抗生素环境风险和生物修复研究中尚存在的问题，展望了未来的研究方向。
关键词：土壤植物土壤动物微生物抗性基因联合修复
Environmental risks of antibiotics in soil and the related bioremediation technologies
Yujie He1,2, Kaiping Zhou1, Yixuan Rao1, Rong Ji1,2
1. State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing 210023, Jiangsu, China;
2. Quanzhou Institute for Environment Protection Industry, Nanjing University, Quanzhou 362000, Fujian, China
Received: June 3, 2021; Accepted: August 18, 2021; Published: September 1, 2021
Supported by: National Natural Science Foundation of China (Nos. 31861133003, 21806075), Science and Technology Bureau of Quanzhou City, China (No. 2019C113)
Corresponding author: Rong Ji. Tel: +86-25-89680581; E-mail: ji@nju.edu.cn.
Abstract: Antibiotics are widely used and prevalently distributed in the environment. The issue of antibiotic resistance genes has posed a huge threat to the global public health. Soil is an important sink of antibiotics in the environment. Antibiotic exposure may introduce adverse effects on soil organisms, and bring indirect but potential risks to human health. Therefore, it is urgent to take actions to remediate antibiotics-contaminated soil. This review summarized effects of antibiotics on phenotype growth of plants, physiological characteristics and community structure of animals, composition and structure of microbial communities, and transmission of antibiotic resistance genes among organisms in soil. Additionally, the potential and prospects of employing antibiotic-resistant soil plants, animals, microorganisms, and their combinations to treat antibiotics-contaminated soil were illustrated. Last but not least, the unaddressed issues in this area were proposed, which may provide insights into relevant research directions in the future.
Keywords: soil plantssoil animalsmicroorganismsantibiotic resistance genescombined remediation
自1928年青霉素被发现以来，抗生素被广泛用于治疗人类和动物的各种疾病，以及促进动物的生长等。然而，人类和动物服用的抗生素约有59%不能被吸收利用，进而随尿液和粪便排放到环境中[1]。抗生素的主要作用机理为攻击细菌的核糖体、抑制细胞壁的合成、破坏类脂膜的完整性，以及阻断碳代谢和DNA的合成[2]。进入到环境中的抗生素会对环境微生物产生负面影响、加速抗生素抗性基因(Antibiotic resistance genes，ARGs) 的传播，并引发一系列免疫学反应等[3]。2014年世界卫生组织就114个国家ARGs的调研结果发表声明称，ARGs问题已对公共卫生构成重大威胁[4]。据报道，全球每年有70万人死于抗生素耐药性，如不采取有效措施，年均死亡人数到2050年将攀升至1 000万人次[5-6]。
土壤是环境中抗生素的重要汇。已有大量研究表明，抗生素会对土壤植物、动物及微生物造成环境风险，包括影响植物的生长，改变土壤动物的生理特性，给微生物带来选择压力，诱发抗性基因在微生物、土壤植物和动物间的产生、表达和传递[7-8]。土壤中的抗生素会对植物、土壤动物和微生物带来危害，但同时利用耐受的生物可以实现对土壤中抗生素的去除。生物修复是利用植物、动物或微生物等直接降解或富集污染物以便进一步处置，从而达到清除或治理污染的目的。相较于物理化学方法，利用生物方法修复抗生素污染土壤更加低耗经济，且可避免二次污染问题[9-10]。本文系统综述了近年来土壤中抗生素污染的环境风险及生物去除技术，分析总结了研究中尚存在的问题，为该领域的研究发展及生物修复应用提供参考。
1 土壤中的抗生素我国已成为发展中国家中仅次于印度的居民用抗生素消耗大国[11]，兽用抗生素消耗量也位居全球之首[12]。据报道，我国近70%的处方药皆为抗生素，而发达国家的占比仅为30%[13]。集中式动物饲养模式的发展大幅度刺激了兽用抗生素的市场需求，我国近半数的抗生素皆被用于动物饲养[14]。统计表明，我国居民及兽用抗生素的主要类别为β内酰胺类、大环内酯类、磺胺类、喹诺酮类和四环素类抗生素[6]。
抗生素可通过制药工厂、医院、居民区和养殖场产生的污水和固体废弃物等多种途径进入环境中，其中土壤是其在环境中的重要汇[15-16]。针对全球抗生素用量排名较高的国家，本文总结了各国土壤中5种主要类别抗生素的检出浓度(表 1)。除巴西和我国土壤中分别检出mg/kg水平的氟喹诺酮类和β内酰胺类抗生素外，其余抗生素的检出浓度皆分布在μg/kg水平。此外，我国****检出或关注的抗生素种类更多。Lyu等[6]对我国土壤中抗生素的溯源结果表明，检出率较高的19种抗生素中，有12种抗生素主要来源于畜禽和水产养殖业的废水与粪便的排放，其余来源于生活污水和工业废水的排放。每年我国会产生超过30亿t的畜禽粪便，其中大部分仅经过简易预处理后即被施用于农田用作肥料[14]，成为土壤抗生素暴露风险的重要来源。
表 1 抗生素用量较高的国家土壤中抗生素污染物的检出浓度Table 1 The detected concentrations of antibiotics in soil in different countries
| Income level | Country | Categories | Antibiotics (concentration, μg/kg) |
| High- income countries | USA | Macrolides | Tylosin (2.4), agricultural soil[17]; tylosin (32), agricultural soil[18] |
| Sulfonamides | Sulfamethoxazole (6.2), sulfadimidine (36), dairy farm soil[19]; sulfadimidine (72.6), agricultural soil[17] | ||
| Tetracyclines | Tetracycline (105), oxytetracycline (25), dairy farm soil[19]; chlorotetracycline (63), agricultural soil[18]; oxytetracycline (9), tetracycline (10), field soil[20]; chlorotetracycline (0.9), agricultural soil[17] | ||
| France | Sulfonamides | Sulfamethoxazole (2.5), soil receiving urban wastewater[21]; sulfamethoxazole (18), subcatchment soil[22]; sulfamethoxazole (nd), sulfadimidine (nd), agricultural soil[23] | |
| Quinolones | Norfloxacin (9.4), ciprofloxacin (8.7), ofloxacin (8.6), agricultural soil[23]; oxolinic acid (5.9), nalidixic acid (22.1), flumequine (6.9), soil receiving urban wastewater[21] | ||
| Tetracyclines | Chlorotetracycline (nd), doxycycline (nd), agricultural soil[23] | ||
| Italy | Macrolides | Tylosin (nd), calf manured soil[24] | |
| Quinolones | Flumequine (77.7), agricultural soil[25] | ||
| Tetracyclines | Oxytetracycline (7), calf manured soil[24] | ||
| Germany | Macrolides | Tylosin (nd), agricultural soil[26] | |
| Sulfonamides | Sulfadimidine (2), livestock farm soil[27]; 4-hydroxysulfadiazine (2 300), sulfadiazine (5.8), agricultural soil[28] | ||
| Tetracyclines | Tetracycline (198.7), chlorotetracycline (7.3), oxytetracycline (nd), agricultural soil[26]; tetracycline (23.5), agricultural soil[28]; tetracycline (295), chlorotetracycline (39), livestock farm soil[27] | ||
| Upper- middle- income countries | Brazil | Quinolones | Enrofloxacin (46.4-6 138), ciprofloxacin (12.6-1 371), agricultural soil[29]; enrofloxacin (17.4-26.7), agricultural soil[30] |
| Turkey | Sulfonamides | Sulfathiazole (nd-400), sulfamethoxazole (nd-110), agricultural soil[31] | |
| Quinolones | Enrofloxacin (nd-50), agricultural soil[31] | ||
| Tetracyclines | Oxytetracycline (nd-500), chlorotetracycline (nd-100), agricultural soil[31]; oxytetracycline (500), agricultural soil amended by poultry manure[32] | ||
| Low- and lower- middle- income countries | India | Quinolones | Ciprofloxacin (3.8-1 900), norfloxacin (< 11), ofloxacin (1.2-19), soil collected from villages[33] |
| China | β-lactams | Penicillin (nd-1 880), soil collected from cattle-producing area[34] | |
| Macrolides | Spiramycin (0.02), tylosintartrate (0.003), anhydroerythromycin (0.06), roxithromycin (0.05), urban surface soil[35]; erythromycin (nd), roxithromycin (nd-0.6), agricultural soil[36] | ||
| Sulfonamides | Sulfamethoxazole (0.6), sulfapyridine (0.01), sulfamethazine (0.05), sulfadiazine (0.3), sulfamonomethoxine (0.01), urban surface soil[35]; sulfadiazine (0.04-0.10), sulfadimidine (0.05-0.40), sulfametoxydiazine (nd-0.08), sulfadimethoxine (nd 0.50), sulfamethoxazole (nd-0.06), agricultural soil[36]; sulfadiazine (1.9-760.1), sulfamethazine (11.1-311.3), sulfadimidine | ||
| Low- and lower- middle- income countries | China | Sulfonamides | (2.6-11.5), sulfamethoxazole (6.4-671.5), urban soil[37]; sulfamethoxazole (0.03-0.9), sulfamethoxine (nd-9.1), livestock farm soil and agricultural soil[38] |
| Quinolones | Norfloxacin (94.6), ciprofloxacin (36.6), enrofloxacin (0.4), fleroxacin (2.4), ofloxacin (6), lomefloxacin hydrochloride (2.3), sarafloxacin (0.2), urban surface soil[35]; norfloxacin (4.3-13), ciprofloxacin (2.9-23), enrofloxacin (6.3-47), lomefloxacin hydrochloride (1.2-2.3), agricultural soil[36]; ofloxacin (0.6-1.6), pefloxacin (nd), ciprofloxacin (0.8-30.1), livestock farm soil and agricultural soil[38] | ||
| Tetracyclines | Tetracycline (2.6-5.2), oxytetracycline (13-80), chlortetracycline (3.9-17), agricultural soil[36]; oxytetracycline (17.6-1 398.5), tetracycline (29.5-976.2), chlortetracycline (8.3-1 590.2), doxycycline (11.1-870.5), urban soil[37]; oxytetracycline (124-2 683), tetracycline (2.5-105), chlortetracycline (nd-1 079), livestock farm soil and agricultural soil[38] | ||
| Notes: the countries listed are the largest consumers of human and animal antibiotics at high-income, upper-middle-income, low- and lower-middle-income levels[11-12]. “nd”: not detected. | |||
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进入到土壤中的抗生素，会和土壤组分及其中的微生物发生吸附和降解等反应，甚至会被土壤中的植物吸收并进一步富集[39-40]。这些反应过程与抗生素自身的性质、土壤组分、土壤环境条件及植物种类等密切相关[8, 40]。例如，相较于其他类抗生素，β内酰胺类抗生素易被水解和生物降解，其在含水率较高的土壤中(28%-37%) 的半衰期为几天甚至是几个小时[41]。高含量腐殖酸的黏土(10 mg/L，7%吸附) 相比于低腐殖酸含量时(1 mg/L，35%吸附) 更不利于土霉素的吸附，这可能是由于pH 11条件下腐殖酸脱氢后的羧基和酚类官能团与呈阴离子状态的土霉素互斥所致[42]。在pH 2.3-7.7的壤砂土、砂壤土和壤土中，两性磺胺二甲嘧啶和磺胺噻唑的分配系数皆呈现递减的趋势，这与其在该pH范围内由阳离子、中性离子到阴离子转变的解离状态相关[43]。就植物吸收富集作用而言，辛醇-水分配系数对数值(log Kow) 在0.5-3.5范围内的抗生素较易被植物吸收[44]，疏水性强的抗生素在植物根系中的生物富集系数(Bioconcentration factor，BCF) 在一定程度上与根系的脂质含量呈正相关关系[45]。共同暴露于含9种抗生素的土壤中，叶菜类作物(生菜) 和根块类作物(胡萝卜) 对抗生素的富集系数存在显著差异[46]。
大量研究表明，尽管抗生素的物理化学性质多样，大部分抗生素在土壤中皆难以被降解，会持久地赋存于土壤中[39]，主要以不可提取态残留的形式存在[40]。虽然该抗生素残留与土壤组分紧密结合，但其被证实仍具有生物活性，且可能会影响土壤中耐药菌的演替过程[47]，造成环境风险。
2 土壤中抗生素的环境风险文献调研表明，截至目前经Web of Science检索的英文期刊文献中，研究或提及土壤中抗生素环境风险的文献有2 487篇；经CNKI检索的期刊文献中，相关的中文文献仅有450篇，如图 1所示。其中，近10年相关英文及中文文献占比分别为64%和73%；近5年二者占比皆为38%。因此，尽管抗生素的使用历史已近百年，但有关其在土壤中环境风险的研究尚在不断探索中。抗生素暴露除了会对土壤生物带来环境风险外，还有可能通过淋溶作用迁移至地下水和地表水中[40]。本文主要就抗生素对土壤生物的影响展开详细讨论。
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| 图 1 国内外研究或提及土壤中抗生素环境风险的文献数目(Web of Science数据库中检索的关键词分别为：soil (title) & antibiotic (topic) & “environmental risk” (topic)；soil (topic) & antibiotic (topic) & “plant growth” (topic)；soil (title) & antibiotic (topic) & effect on organism (topic)；soil (title) & antibiotic (topic) & effect on microorganism (topic)；soil (title) & antibiotic resistance gene (topic). CNKI数据库中检索的关键信息(篇名、关键词及摘要)分别为：土壤 & 抗生素 & 环境效应；土壤 & 抗生素 & 植物 & 影响；土壤 & 抗生素 & 微生物 & 影响；土壤动物 & 抗生素 & 影响；土壤 & 抗性基因. 不同关键词检索所得的重复文献已作合并处理) Fig. 1 The amount of studies that investigated or referred environmental risks of antibiotics in soil. The keywords used in Web of Science including soil (title) & antibiotic (topic) & "environmental risk" (topic); soil (topic) & antibiotic (topic) & "plant growth" (topic); soil (title) & antibiotic (topic) & effect on organism (topic); soil (title) & antibiotic (topic) & effect on microorganism (topic); soil (title) & antibiotic resistance gene (topic). The searching information (title, keywords, and abstract) used in CNKI including soil & antibiotic & environmental risk; soil & antibiotic & plant & effect; soil animal & antibiotic & effect; soil & antibiotic & microorganism & effect; soil & antibiotic resistance genes. Overlapped studies that retrieved according to different terms have been merged. |
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2.1 对植物的影响土壤中抗生素对植物的效应研究多关注植物的表型生理指标，包括萌芽、生物量、根茎叶的发育、营养元素及光合色素含量等。例如，鲍艳宇等[48]发现当土壤中四环素和土霉素的投加浓度超过1 000 mg/kg时才会影响小麦种子发芽，因此萌芽率可能不宜作为抗生素生态毒理学的敏感指标。磺胺甲噁唑和磺胺二甲嘧啶对水稻株高的半数有效浓度(Median effective concentration，EC50) 分别为38和220 mg/kg；对根长的EC50分别为13和43 mg/kg[49]。土壤中10 mg/kg的氧氟沙星、阿莫西林、左氧氟沙星、环丙沙星和氨苄青霉素皆会抑制水稻根和茎的生长及其营养元素的含量，其中氧氟沙星对根重、茎重、根长及茎长的抑制率最高，分别为67%、66%、53%和32%；氨苄青霉素对铁和总碳含量的抑制率最高，分别为33%和41%；而左氧氟沙星对蛋白质含量的影响最大，相较于空白对照组降低了30%[50]。由于对植物体内叶酸的抑制作用，10 mg/kg磺胺嘧啶即显著抑制了(> 50%) 10种湿地植物光和色素的合成[51]。Mukhtar等[50]亦通过彗星电泳实验研究了5种抗生素对水稻根细胞的基因毒性，结果发现左氧氟沙星(10 mg/kg) 对细胞DNA链造成的断裂效应最为明显，损伤率高达98%。
相较土壤实验而言，更多的评估抗生素对植物影响的实验是在水培体系中完成的。目前，已有多篇综述总结了水培体系中抗生素对作物及非作物的表型生理指标带来的负面效应[8, 39, 52-54]，其中抗生素对植物造成负面效应的EC50值多分布在几μg/L到几mg/L的水平。例如，Opri?等[55]就9种抗生素(0.5 mg/L) 对小麦光合作用的影响作出了实验研究，结果表明头孢曲松和环丙沙星显著抑制了小麦总叶绿素的合成；强力霉素和氨苄青霉素显著降低了光合电子传递速率。此外，水培实验中抗生素暴露被证实会引发植物的应激机制。强力霉素(0.5-25.0 mg/L) 会引发拟南芥的线粒体蛋白应激毒性，以及核基因表达和线粒体功能的改变[56]。涵盖5个抗生素类别的9种抗生素(0.5 mg/L) 皆会诱导小麦叶片释放单萜和绿叶挥发物(脂氧合酶产物)，而这两种次生挥发性代谢物是植物应对外源刺激的重要标志物[55]。类似地，氟喹诺酮类抗生素(5 mg/L) 被证实会引发小麦的氧化应激，具体表现为促进抗氧化物总酚的生成以及抑制过氧化氢酶和过氧化物酶的活性，且这两种酶活的抑制效应随抗生素暴露剂量的增大而愈发明显，说明抗生素的暴露已对细胞膜造成不可逆的破坏，超出了这些抗氧化酶的协调能力[57]。
农田土壤中施用含抗生素的粪肥除了会对作物带来直接影响外，还可能导致抗生素通过作物富集经食物链传递到人体内带来潜在危害。已有的关于作物富集抗生素的研究或在水培体系，或在温室和农田条件下开展，实验设置差异较大，导致据此推测的人体暴露抗生素的风险差异较大：人体年暴露抗生素剂量范围为0.01 μg/g (据芹菜叶中磺胺类抗生素富集量推算)到8 456 μg/g (据稻谷中的氯霉素富集量推算)，远低于抗生素每日20-200 mg的最低治疗剂量[52]。然而亦有少数研究发现，抗生素极端暴露条件下作物富集作用会对人体健康带来潜在风险。例如，暴露于500 mg/L的四环素中20 d后，生姜的可食部分会富集28.1 mg/kg的四环素，按我国日均摄入生姜量150 g/d计，儿童(按30 kg计) 食用该生姜即会超出四环素的日允许摄入剂量30 μg/(kg·d)[58]。
2.2 对土壤动物的影响进入土壤中的抗生素除可被植物吸收富集外，亦可被土壤动物富集[59]，进而影响其生理特性。研究表明，当土壤中拉沙里菌素的浓度大于4.5 mg/kg时，鼠妇虫Porcellio scaber即表现出对污染土壤明显的回避行为[60]；慢性毒性测试表明，土壤中莫能菌素对白符跳虫Folsomia candida幼虫产率的EC50值为96 mg/kg[61]。蚯蚓是土壤动物最重要的类群之一，可参与土壤中有机质分解和营养元素的循环，影响土壤结构和功能[62]。土壤中莫能菌素对暴露14 d的赤子爱胜蚓Eisenia fetida的致死率EC50为346 mg/kg[63]；暴露于30 mg/kg的强力霉素56 d后，赤子爱胜蚓的幼蚓产率被抑制了62%[64]。相较于繁殖速率，?i?ek等[60]发现以蚯蚓的回避行为表征抗生素的影响更为灵敏，拉沙里菌素对赤子爱胜蚓繁殖速率和回避行为的EC50分别为69.6和12.3 mg/kg。除影响蚯蚓的生理特征外，抗生素还会对其DNA和肠道微生物的组成造成负面影响。彗星电泳实验结果表明，四环素和金霉素暴露浓度为0.3 mg/kg时，即会对暴露7 d的赤子爱胜蚓DNA造成损伤[65]。赤子爱胜蚓暴露于0.1-10.0 mg/kg磺胺甲噁唑中28 d后，其肠道内微生物群落的多样性显著降低，群落组成发生紊乱[66]。
抗生素暴露不仅会对单一土壤动物的生长造成影响，也会对土壤动物群落层面带来负面效应。研究表明，向3块10 m2的黑土农田中分别喷洒1.6、9.6、32.4 mL的甲氨基阿维菌素苯甲酸盐原药溶液后，土壤动物群落(涵盖4纲11目共31个类群) 的个体密度和类群数量显著降低，且该降低趋势随原药喷洒浓度的增高而增强；且群落的垂直分布格局发生改变，表聚性特征被削弱，土壤动物更多地聚集在5-15 cm土层[67]。事实上，除对直接吸收富集抗生素的土壤动物造成直接危害外，抗生素暴露还有可能通过食物链传递对这些土壤动物的捕食者造成间接危害。然而，目前涉及土壤动物经食物链传播有机污染物风险的研究多集中于对全氟烷基污染物、卤代阻燃剂和农药等的研究[59]，鲜见对抗生素的相关报道。
2.3 对土壤微生物的影响由于大多抗生素都有较宽的抗菌谱，可抑制多种细菌和真菌，因此势必会对土壤中的微生物群落结构和功能造成影响，甚至会破坏土壤生态系统功能的稳定性[68-69]。截至目前，已有多篇文献研究了抗生素对土壤微生物活性、多样性、群落组成和功能等的影响[70-71]。向土壤中施加磺胺甲噁唑、磺胺二甲嘧啶和甲氧苄啶2 d后，即会显著影响土壤的呼吸作用，三者的EC10分别为7、13、20 mg/kg[49]。在水稻土中施加100 mg/kg的磺胺甲噁唑21 d后，会显著抑制微生物对31种碳源的利用率及降低其香农指数[72]。Westergaard等[73]的研究指出，相较于多样性而言，微生物群落结构用于评估抗生素对土壤微生物的影响更为合理，其发现土壤暴露于泰乐菌素(2 000 mg/kg) 的2个月中，微生物多样性仅发生了短暂的变化，而泰乐菌素对群落结构的变化却是持久性的。类似的，暴露于1-10 mg/kg的磺胺嘧啶48 d后，土壤中原先受抑制的微生物多样性也得以恢复[74]。
基于细菌和真菌的C/N比以及革兰氏阳性菌和阴性菌细胞壁结构的差异[69, 75]，土壤中细菌/真菌比及革兰氏阳性菌/阴性菌比常被用于反映微生物的群落结构，表征土壤生态系统的功能[69]。向壤砂土和粉砂壤土施用含额外添加10 mg/kg磺胺嘧啶的粪肥，4 d后土壤微生物的组成皆向真菌演变，细菌/真菌比分别降低了约76%和77%；粉砂壤土中的革兰氏阳性/阴性菌比增加了约37%[76]。此外，一些土壤微生物的代谢功能会受到抗生素暴露的显著影响。DeVries等[77]总结了多种抗生素对土壤氮循环的影响，结果表明抗生素暴露主要体现为抑制土壤微生物的硝化和反硝化作用。然而，也有少量报道表明，低浓度环丙沙星和诺氟沙星(1 mg/kg) 暴露促进了土壤微生物的硝化作用[78-79]；500 μg/kg的磺胺甲噁唑、磺胺嘧啶、甲基盐霉素和庆大霉素显著抑制土壤的反硝化作用，但当暴露浓度低于1 μg/kg水平时反而促进了反硝化过程[80]。因此，抗生素暴露对土壤氮循环的影响依其种类及暴露浓度而定。抗生素还被证实会影响土壤的产甲烷及铁还原等过程：泥炭土中的微生物经500 μg/kg的磺胺甲噁唑暴露后产甲烷速率得以促进[81]；1 mg/kg的土霉素和磺胺甲噁唑暴露即会抑制农田土中96%和92%的三价铁的还原效率[82]。
抗生素暴露除了会对土壤微生物带来上述直接影响外，还会间接对微生物产生抗性选择压，且该间接影响一般在ng/kg水平上即会发生，主要体现为诱导抗性基因ARGs和抗性菌的形成[70]。此外，目前针对抗生素对土壤微生物影响的研究中，所设抗生素的暴露水平多高于其在土壤中的环境浓度。参考培养基体系中金黄色葡萄球菌的相关研究，多种抗生素在亚致死浓度水平上(低于mg/L) 可调控金葡菌关键基因的功能，促进被膜的形成及其活性[83]。
2.4 抗性基因的传播自然环境中的细菌和真菌是最初抗生素的主要来源，也是抗生素耐药性的发源地，人类活动的干扰(如抗生素的大量使用及人工合成等) 引发了自然环境生态系统中抗性基因的进化和传播[84-86]。微生物可通过多种途径对抗生素产生抗性，包括生成使抗生素失活的酶、改变细胞壁的通透性阻止抗生素进入、改变细胞内抗生素的作用靶位、激活细胞壁的外排泵运输系统阻碍抗生素在细胞内的蓄积等[2, 70]。抗生素的耐药性不局限于能产生抗生素或抗生素抗性的耐药菌体内，耐药菌还会通过可移动基因元件(Mobile genetic elements，MGEs) 将携带的ARGs经接合、转导和转移过程传递给其他生物[14, 87]。目前，尚不清楚环境中抗生素暴露与ARGs传播之间是否存在必然的联系，既有研究支持抗生素的存在促进了ARGs的传播，也有研究表明两者无直接的联系[69]。土壤中抗性基因的传播是受多因素影响的过程，但不可否认的是，抗生素暴露是其中相当重要的一个原因。
短期来看，无论是实验室小试实验[88] (63 d)还是野外实验[89] (132 d)，将额外添加磺胺嘧啶的猪粪(30 mg/kg) 施用于种植玉米的土壤中，磺胺类抗性基因sul1和sul2的相对丰度较不添加磺胺嘧啶的粪肥对照组显著增高，且由于根际土壤中磺胺嘧啶的检出浓度较非根际土低，因此根际土中两种ARGs的相对丰度也更低。长期来看，因抗生素暴露引发的ARGs随暴露时间的递增或维持在较高的水平，或回归到暴露前的状态。向野外农田土中施用含抗生素的粪肥后，Hong等[90]发现土壤中两种四环素类抗性基因tetQ和tetZ，以及整合酶基因intI1和intI2的丰度增长了6倍多，且该趋势一直持续到16个月实验结束；Marti等[91]却发现土壤中3种抗性基因sul1、erm(B) 和str(B) 以及整合酶基因intI1的丰度仅在实验初始的几周有所上升，在随后1年的时间内回归到初始水平。然而，从更长的时间维度来看，长期使用抗生素的确增加了土壤中ARGs的量。Knapp等[7]收集了荷兰5个地区跨度长达68年的施用过粪肥和无机肥的农田土壤样品(1940-2008年)，测定其中18种ARGs的相对丰度变化，结果表明所有ARGs的丰度皆呈递增趋势，尤其是四环素类ARGs的增幅自1970年至2008年上涨了15倍之多。
受抗生素暴露后，土壤中除微生物的ARGs丰度会有所改变外，土壤植物和动物中的ARGs水平也会受影响。向种植生菜的农田土中施加含抗生素的粪肥65 d后，土壤微生物、生菜及蚯蚓中152种ARGs的整体水平皆显著增高[92]。生姜暴露于500 mg/L的四环素中20 d后，其可食部分即有四环素ARGs及intI1检出[58]。尽管目前还没有直接的证据表明ARGs可以从土壤环境转移到人类病原体中[93]，但已有研究指出，食用在抗生素污染的土壤中种植的蔬菜是一种潜在暴露于土壤ARGs的途径[14, 94-95]。总之，不可否认的是，抗生素及ARGs污染在一定程度上增加了人类病原体经MGEs途径产生耐药性的风险[96]。
3 抗生素污染土壤的生物修复与土壤中抗生素的环境风险研究相比，截至目前研究或提及抗生素污染土壤生物修复的英文文献仅有91篇(以soil为标题、antibiotic和bioremediation为主题在Web of Science中检索的英文期刊文献数)；而相关的中文文献为19篇(以土壤、抗生素和生物修复为“篇关摘”在CNKI中检索的中文期刊文献数)，表明对抗生素污染土壤生物修复的研究尚处于初探阶段。
3.1 植物修复截至目前，土壤中的植物修复多被用于治理重金属和多环芳烃等污染，较少研究关注该修复技术对抗生素的去除效应[8, 39]。已有涉及植物修复土壤中抗生素污染的研究仍处于初步实验阶段。柳树因其生长快、根系发达以及可被用作可再生能源等特点，常被用于修复重金属和多环芳烃等污染的土壤[97]。Michelini等[97]基于水培实验证实了柳树对磺胺类抗生素的去除潜力，爆竹柳Salix fragilis根部在25 d内对620 mg/L的磺胺二甲氧嗪的吸收量为27.4 mg/g。草本植物中，利用芥菜Brassica juncea吸收225 mg/L的四环素，1个月后水培体系中四环素仅有7%残存[98]。基于芥菜的修复潜力，Cui等[99]间作种植芥菜和黑麦草Lolium multiflorum，考察了其对抗生素、ARGs和重金属复合污染土壤的修复效果。结果表明，针对8种抗生素污染(35-432 μg/kg) 的土壤，间作种植对抗生素的去除效率为5%-100%，相较于单独种植芥菜和黑麦草提高了2%-54%；间作种植条件下，植物对抗生素的BCF值以及对ARGs的去除效率相较于单独种植分别提高了1.2-17.8倍及11%-23%。同样是针对抗生素和重金属复合污染的土壤，周显勇等[100]利用重金属超积累植物伴矿景天Sedum plumbizincicola对11种抗生素污染(0.6-12.7 μg/kg) 的菜地土壤进行为期9个月的修复，实现了对抗生素36%-76%的去除。
尽管实验结果显示植物具有修复抗生素污染土壤的潜能，但植物修复技术的实际应用仍面临着挑战。参照植物对重金属和多环芳烃等污染土壤的修复，大量的实验室和温室实验以及一些野外实验皆证实了植物修复的可行性[101]，然而实际应用过程中植物需应对气候条件变化、病原菌和食草动物的胁迫以及与其他植物的竞争等不确定因素[102]，这些因素可能会对植物修复的效率造成不可忽视的影响。因此，为在实际应用中最大程度发挥植物修复的效益，需采取措施尽可能地减小植物在修复过程中面临的各种胁迫(详见下文3.4部分)。
3.2 土壤动物修复作为生物量最大的土壤大型动物，蚯蚓具有发达的肠道系统，消化能力强且肠道内富含微生物[103]，其被证实可有效地修复重金属和有机物(如多环芳烃、多氯联苯、农药等) 污染的土壤[104-107]。相较重金属和多环芳烃等有机污染物而言，仅有少量研究报道了蚯蚓对抗生素污染土壤的修复能力。两种生态型蚯蚓赤子爱胜蚓(表层种) 和壮伟远盲蚓(Amynthas robustus，表层种) 40 d内对灭活砖红壤中5 mg/kg四环素的去除效率分别为26%和23% (已扣除灭活土壤吸附)；其在活性土壤中对四环素的去除效率分别为64%和75%，相较未添加蚯蚓的活性土壤去除率分别提高了37%和50%[108]。该研究中蚯蚓肠道和蚓粪中即包含可降解四环素的微生物，且蚯蚓的添加可中和土壤pH、消耗土壤有机质、促进土壤微生物对四环素的降解转化。类似的，蚯蚓被证实会促进土壤微生物对土霉素的降解转化，可达到修复的目的。研究表明，赤子爱胜蚓或可通过增加土壤微生物总量及特定降解微生物的丰度以促进对土霉素的降解[109]；或可通过增加土壤微生物的活性和多样性以加速对土霉素及其两种代谢产物的转化[106]。
上述利用蚯蚓修复抗生素污染土壤的研究中，蚯蚓主要是通过影响土壤性质及微生物活性而非自身的富集作用来促进抗生素转化，且所涉土壤中抗生素的暴露浓度多低于几mg/kg水平。该暴露水平下，蚯蚓的生理特征不会发生明显改变，但其DNA和肠道微生物可能会遭受负面影响(如2.2部分所述)。然而，目前尚未有报道明确蚯蚓在强化抗生素去除的同时，其自身是否会受抗生素长期暴露的影响。
3.3 土壤微生物修复目前已有众多****从污水、污泥、粪肥和土壤等介质中筛选出以降解四环素类和磺胺类抗生素为主的特异性降解菌[110]，例如成洁等[111]从长期与鸡粪接触的土壤中筛选出的可降解土霉素、四环素和金霉素的木糖氧化无色杆菌Achromobacter xylosoxidans及枯草芽孢杆菌Bacillus subtilis。从土壤中筛选出的降解菌多在培养基体系中被验证可实现对抗生素的高效降解，如Leng等[112]从牧场土中筛选出四环素降解菌Stenotrophomonas maltophilia strain DT1，4 d内可将30 mg/L四环素完全转化为代谢产物。然而，少有研究将所筛得的降解菌应用于土壤中微生物强化去除抗生素。
Shi等[113]从活性污泥中筛得可降解四环素和土霉素的Arthrobacter nicotianae OTC-16，以10% (V/W) 的比例接种到pH为8.36的碱性土壤中，该降解菌所在的贝氏谷氨酸杆菌属Glutamicibacter成为土壤中的优势菌属，可于10 d内去除74%浓度为100? mg/?kg的土霉素，相较于不添加降解菌的对照组去除效率增高46%，且显著降低了土壤中ARGs的丰度。为促进降解菌在土壤中的定殖效果，提高其微生物的强化效率，Hong等[114]利用甘蔗渣的多孔、机械强度高、易负载、易生物降解及可被微生物用作碳源和营养源等优点，将从猪粪中分离所得的土霉素降解菌Burkholderia cepacia负载在甘蔗渣上，其30 d后对100 mg/kg土霉素的去除效率达到79%，较未负载的游离菌高19%；此外作者还对负载菌的接种剂量以及修复环境的温度和pH等条件作出了优化。Hirth等[115]还对比了土壤修复中纯降解菌和菌群对磺胺二甲嘧啶的去除效率。作者从长期受纳抗生素的农田土中筛选得到磺胺二甲嘧啶的降解菌Microbacterium sp. strain C448[116]及包含该降解菌的菌群，进一步将其负载在2–4 mm的黏土颗粒上并投加到土壤中进行修复(10颗每50 g土)，49 d后负载纯菌及菌群对1 mg/kg磺胺二甲嘧啶的矿化效率分别为39%和8%；112 d后负载菌群(12颗每35 g土) 对1 mg/kg磺胺二甲嘧啶的矿化率高达51%，剩余量中43%以不可提取态残留的形式赋存在土壤中。目前，已有一些专利提出将降解菌以固化小球的负载形式[117]或设计降解菌群[118]用作土壤修复剂治理磺胺类及四环素类抗生素的污染。
同植物修复一样，将特异性降解菌或菌群实际应用于土壤中强化去除抗生素仍存在很多不确定的因素，主要体现在接种菌在受试场地中定殖后，其活性和降解能力受环境因素如温度、土壤pH、含水率、有机质含量等以及与土著微生物生态位竞争的影响[119]。因此，需要采取措施提高接种微生物的定殖能力，包括：将微生物负载在生物炭、纳米材料等载体上[119]，基于分子生物学分析手段挖掘降解菌的关键基因元件和降解酶等[110]，向微生物中导入具有降解抗生素功能的基因[120]，或利用合成生物学技术构建具备彻底代谢目标抗生素的工程菌[121]。
3.4 土壤生物联合修复植物、土壤动物和微生物三者在土壤生态系统中相互作用、密不可分，因此可以借助其互作关系联合修复污染土壤(图 2)。土壤动物如蚯蚓其排泄物及脱落物可为植物提供养分，同时蚯蚓也可利用植物的残枝落叶生长[122]。Wu等[103]的研究表明，在种植香根草的土壤中添加蚯蚓，其根部对镉(20 mg/kg) 的富集效率提高约57%，该促进作用主要是由于蚯蚓改变了土壤的物理化学性质所致。植物根系分泌代谢物和氧气，可为根际微生物提供碳源、能源和有氧条件；反之根际微生物能为植物提供营养元素、合成可减轻植物胁迫的物质以及保护植物抵御病虫害等[101]。研究证实，利用植物-植物促生菌、植物-污染物降解菌以及植物-植物内生菌的协同作用可有效修复土壤中的重金属、多环芳烃和多氯联苯的污染[101, 123-124]。例如，种植黑麦草Lolium multiflorum的土壤中添加多环芳烃的降解菌Acinetobacter sp. (3.3×106 CFU/g) 后，其对土壤中50 mg/kg芘的去除率在60 d内可达约95%，相较于单独利用黑麦草(约88%) 和降解菌(约62%) 对芘的去除率更高[125]。然而，目前还鲜有研究利用互作的土壤生物联合修复土壤中的抗生素污染，该方向有待进一步的展开。
 |
| 图 2 土壤中抗生素的生物修复(包括植物、土壤动物、土壤微生物及其联合修复. 植物和土壤动物可通过分泌代谢物和氧气、改变土壤的理化性质等途径影响土壤微生物群落的结构与功能) Fig. 2 Bioremediation of antibiotics in soil utilizing plants, soil animals, microbes and their combination. Plants and soil animals are capable of altering the compositions and functions of soil microbial communities by releasing exudates and oxygen, and changing physiochemical properties of soil, respectively. |
| 图选项 |
4 总结与展望土壤是环境中抗生素非常重要的汇，且大部分抗生素在土壤中皆难以被降解，进而会持久性地赋存于土壤中，并对其中的植物、动物和微生物造成环境风险，甚至在极端暴露条件下会通过食物链进入到人体内对其健康带来潜在危害。此外，抗生素暴露会增大土壤中植物、动物和微生物中ARGs的丰度，该问题更是不容忽视，其在一定程度上增加了人类病原体经MGEs途径产生耐药性的风险。生物修复方面，已有研究充分证实了植物、土壤动物和微生物修复抗生素污染土壤的潜力，尤其是参照土壤中其他有机污染物的修复现况，利用土壤生物联合修复方法治理抗生素污染更具应用前景。
然而，目前有关土壤中抗生素环境风险的研究尚存在一些不足之处：1) 土壤中抗生素对植物的效应研究多关注植物的表型生理指标，关于抗生素对植物光合作用及应激机制影响的研究更多的是基于水培实验得出的结论。土壤组分和微生物对抗生素的吸附和降解等作用会导致其生物可及性和生物可利用性降低，因此水培实验所得出的结论在一定程度上高估了土壤中抗生素对植物造成的环境风险。2) 抗生素对土壤动物的影响研究多以致死率和繁殖速率为评估指标，对动物DNA的损伤等更为灵敏的指标研究较少。此外，已有关于抗生素暴露对蚯蚓影响的研究主要针对食腐型赤子爱胜蚓，鲜有对食土型蚯蚓的研究。3) 目前土壤中抗生素的风险研究多基于mg/kg暴露水平，高于土壤的实际暴露水平(表 1，μg/kg)。参考其他有机污染物，环境浓度下其暴露可能会在分子水平上对植物和微生物的代谢产生影响[126-127]，因此有必要更进一步研究环境浓度下抗生素对土壤生物如在分子水平上的影响。4) 抗生素对土壤微生物多样性等的影响可能只是短暂非持久性的，然而已有研究多关注短期暴露下抗生素对土壤生物的效应，较少致力于探索抗生素的长期暴露效应。5) 已有关于作物富集抗生素的研究中实验设置差异较大，导致据此推测的人体暴露抗生素的风险评估结果差异较大，且目前也鲜有关于土壤动物经食物链传播抗生素风险的报道，因此仍较难准确评估土壤中抗生素污染间接对人体健康带来的潜在风险。
生物修复抗生素污染土壤的研究尚处于起步阶段，且多针对重金属和多环芳烃、多氯联苯、农药等有机污染物。值得注意的是，土壤中抗生素污染的去除有别于其他有机及无机污染物的去除，因为它还涉及ARGs传播风险的防控。相较于物理和化学修复方法，生物修复的一大优势即在于利用自然环境中生存的对抗生素耐受的植物、动物和微生物为载体实现对抗生素的强化去除[85]，在一定程度上可避免由于人为引入的元素加剧ARGs的传播。尽管耐受植物、动物和微生物具备去除抗生素的潜力，但将其拓展到实际应用中仍面临挑战。例如，修复过程中所用强化植物需应对气候条件变化、病原菌和食草动物的胁迫以及与其他植物的竞争等不确定的因素；植物或动物吸收富集抗生素后需进一步处置；强化微生物需克服环境因素和土著微生物生态位竞争的影响；需考虑生物修复周期问题等。未来需运用更先进的分析手段，挖掘土壤生物富集或降解抗生素的关键机制，充分利用土壤生物的互作关系，开发抗生素污染土壤的高效联合修复技术。
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