胡 琴,张利兰,2*,易美玲,杨 锐

芘对土壤微生物氮转化功能菌群的影响特征

胡 琴1,张利兰1,2*,易美玲1,杨 锐1

(1.重庆大学,三峡库区生态环境教育部重点实验室,重庆 400044；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)

通过构建好氧降解微环境,分析环境浓度下的芘(12.09mg/kg)对土壤酶活性,氮转化全过程以及相关功能微生物的影响.结果发现,芘仅在降解第1d显著促进了脲酶活性,而在降解最初和后期均显著刺激了脱氢酶活性.从细菌群落结构分析可知,由于氨氧化菌()相对丰度的变化,导致芘在处理前期对其介导的好氧氨氧化,硝化功能表现为促进作用,在后期表现为抑制作用,而对于固氮细菌(,和),尿素分解细菌()以及硝酸盐还原细菌()则作用相反.与微生物群落结构以及相关功能预测的变化不同,功能基因定量分析表明,芘虽在培养初期对固氮基因H表现为抑制作用,但H的丰度呈增长趋势.结合土壤氨氧化和反硝化过程中关键酶活性及编码基因的变化,芘在培养前期未促进氨氧化过程,但在15d后明显抑制了土壤氨氧化和反硝化过程,其中对氨氧化过程的抑制作用更为显著.本研究阐明了芘对土壤微生物氮转化过程的影响特征,为了解芘的环境风险提供重要参考价值.

芘；土壤酶活性；氮转化细菌群落；氮转化过程

多环芳烃(PAHs)是一类具有“三致”毒性的持久性有机污染物[1],由于其亲脂性和惰性,可以通过生物富集等方式长期存在于生态系统中,使其具有较大的生态环境风险[2-3].土壤是PAHs重要的赋存介质,土壤中PAHs浓度构成主要受到人类活动的影响.在城市,工业和农业土壤中,交通尾气,石油泄漏,燃煤和生物质燃烧通常是造成表层土壤PAHs污染的主要原因,其浓度范围分别为几～几千μg/kg[4-7],十几～几万μg/kg[8-10]和几～几百μg/kg[11-13].其中高分子量多环芳烃(high molecular weight PAHs, HMW-PAHs)为主要成分,其浓度整体上略高于低分子量多环芳烃(low molecular weight PAHs, LMW-PAHs),可能由于HMW-PAHs挥发性差且较难被生物降解利用.土壤是自然界最复杂的生态系统之一,健康的土壤是保证粮食安全的关键.因此,土壤中PAHs的环境行为及生态效应值得广泛关注.

微生物群落作为土壤生态系统的骨架,在维持土壤健康方面发挥重要作用,其结构的多样性和稳定性是保持土壤生态系统稳定性和适应性的基础[14].微生物介导的土壤氮转化是地球氮转化系统的中枢环节,对维持各个圈层的生态稳定具有重要意义[15].在PAHs胁迫下,部分微生物通过降解代谢等途径改变其群落组成或活性,进而影响着土壤中的氮转化.例如,LMW-PAHs的芴和菲对氮矿化细菌表现出毒性效应,且具有抑制土壤硝化潜势的作用,抑制强度与PAHs的生物可利用性呈正相关[16];前期研究发现,环境浓度下的菲(12.21mg/kg)增加了土壤中反硝化和固氮微生物的数量,并显著抑制土壤氨氧化过程[17];相近浓度下HMW-PAHs的苯并[a]芘对反硝化和固氮微生物表现出毒性效应,抑制土壤的好氧氨氧化,固氮以及硝酸盐还原过程[18].由此可知,不同结构的PAHs对土壤氮转化微生物产生的毒性效应具有差异性.

芘作为典型的HMW-PAHs之一,由四个对称的苯环组成,结构与许多致癌的HMW-PAHs相似,比LMWPAHs的“三致”毒性更强[19].由于芘的理化性质和生物可利用度等原因,可通过食物链累积在环境中持续存在,产生一定的生态毒性效应[20-21].因此,芘通常用作研究HMW-PAHs环境行为及生态效应的模型[22-23].已有研究证实,低浓度的芘(1mg/kg)对沉积物中的固氮菌有一定的刺激作用,而浓度较高时(10和100mg/kg)则表现为明显的抑制作用[24].此外,芘(0～500μg/g)会显著改变氨氧化细菌和氨氧化古菌丰度的比值,且浓度越大,对氨氧化微生物的活性影响越大[25].基于目前芘对土壤氮转化全过程以及相关微生物功能结构的变化尚不清晰,本研究选择以环境浓度下的芘作为目标污染物,通过分析其对土壤酶活性,氮转化过程以及相关微生物功能结构随时间变化的影响,揭示其对土壤微生物氮转化功能菌群影响特征,为评估土壤中芘的微生态系统功能提供方法参考.

1 材料与方法

1.1 土壤样品采集

1.2 室内培养实验

本研究使用120mL带丁基橡胶塞的棕色玻璃小瓶构建好氧降解微环境,设置以下3个处理组:(1)向土样中添加有机溶剂所溶解的芘(芘处理组,PYR),测得负载后的芘浓度为12.09mg/kg,此负载浓度接近环境浓度[26-28];(2)向土样中添加等量有机溶剂(空白对照组,CK).(3)向经过高温高压(0.1MPa,121℃)灭菌1h的土壤中负载等量芘(非生物对照组),并将土壤置于室温下24h进行复苏,复苏后用生理盐水重悬土壤并取上清液进行涂板,以确保灭菌后的土壤中无微生物活性.将所有处理后的土壤置于通风橱中,待有机溶剂充分挥发后分装至棕色小瓶.每个瓶中分装10g(干重)土样,分装好后置于22℃恒温培养箱中培养.每个处理组3个重复.土壤含水率保持在田间持水量的50%～60%,培养过程中通过称重监测含水率变化.分别在培养0, 1, 3, 7, 15, 30, 60d后进行破坏性取样,以方便后续的提取和测量.

1.3 土壤芘的提取与测定

土壤中芘的提取采用溶剂超声萃取法[29].以甲醇为提取剂提取土壤样品中的芘:向2g(干重)土样中加入20mL甲醇后进行超声提取(60Hz,50℃)1h,然后以8000r/min离心10min,上清液过0.22μm有机滤膜后使用高效液相色谱仪进行芘浓度的测定.高效液相色谱仪配备SB-C18色谱分离柱和二极管阵列检测器,流动相为甲醇和水(芘90:10),流速为1mL/min,芘检测波长为240nm.超声萃取提取土壤样品中芘的回收率为80～93%.

1.4 酶活性测定方法

土壤酶来源于土壤微生物,植物等,是反映土壤健康和质量的一项重要指标[30-31].其中,土壤脲酶参与尿素水解,其活性反映了土壤氮素代谢的旺盛程度和无机氮供应能力[32].而脱氢酶是反映有机物降解功能的酶,可作为微生物氧化还原活性的重要指标,指示土壤有机污染物的转化速率[33].土壤脲酶活性测定采用S-UE试剂盒(索莱宝,中国),称取过0.0425mm筛的0.05g土壤样品于2mL离心管中,加入20μL甲苯溶液,在室温下放置15min,再加入90μL尿素溶液和190μL缓冲液,37℃下培养24h后离心并加入显色剂,于630nm处测定吸光值表示产生的氨氮含量,以每单位时间每克土产生的氨氮含量表示脲酶活性.土壤脱氢酶活性测定:向土壤中加入5-三苯基氯化四氮唑(TTC)溶液,在37℃培养24h后用甲醇提取产生的三苯甲酰亚胺(TPF),于485nm处测定吸光度.

氨单加氧酶(AMO),硝酸盐还原酶(NAR)和亚硝酸盐还原酶(NIR)作为氮转化酶,其活性反映了土壤中的相关氮转化速率.其中AMO作为催化土壤中NH4+转化为NO2-的关键酶,在土壤硝化过程中发挥重要作用[34].土壤氨单加氧酶(AMO)活性测定主要采用Yang等[35](2020年)的方法:向2g(干重)土样中加入20mL磷酸盐缓冲溶液(PBS, pH7.4),于37℃黑暗震荡培养4h (150r/min)后离心取上清液,测定上清液产生的亚硝酸盐含量.另外,NAR和NIR分别催化土壤中NO3-和NO2-的还原过程,这是反硝化过程的前两个限速步骤.基于硝酸盐和亚硝酸盐还原反应原理,土壤硝酸盐还原酶(NAR)和亚硝酸盐还原酶(NIR)活性采用S-NR和S-NiR试剂盒(索莱宝,中国)测定.

1.5 土壤DNA提取及功能基因定量分析

称取0.25g土壤样品,根据土壤DNA提取试剂盒(Qiagen,美国)操作说明提取土壤微生物DNA.并用超微量分光光度计(Implen,德国)测定DNA浓度与质量.以提取的土壤微生物总DNA为模板,使用荧光定量PCR仪(Bio-Rad,美国)对氮转化功能基因AOAA, AOBA,H,G和S进行定量分析,引物信息参照文献[18].荧光定量PCR反应体系体积为20μL:前/后引物各0.6μL,DNA模板0.8μL,无菌水8.0μL,2×SYBR Green Supermix10μL.各基因反应条件和程序见参考文献[23].

1.6 高通量测序

利用通用引物515F(5-GTGCCAGCMGCC- GCGGTAA-3)和806R(5-GGACTACHVGGGTWT- CTAAT-3)在Illumina MiSeq平台(Majorbio, Shanghai, China)上扩增土壤细菌16S rRNA基因V4区进行测序分析.使用Qiime 2平台对原始扩增子进行处理.首先使用DATA2对16S rRNA特定区域的高通量序列进行降噪处理,以获得每个样本的扩增子序列变体(Amplicon sequence variant, ASV).然后通过Bayes注释方法,采用Silva 13注释数据库对获得的16S rRNA ASV进行分类.为最大程度降低采样深度的影响,将ASV集进行了抽平处理,以进行下游分析.通过质量筛选,共获得了86960个16S rRNA高质量序列,包含19860个细菌ASV.从19860个细菌ASV中筛选出40种与氮转化相关的细菌属.同时,使用FAPROTAX对细菌氮转化功能进行注释[36].

1.7 数据分析

芘在土壤中的好氧微生物降解动力学拟合为零级动力学方程(1).芘的半衰期由方程(2)计算.

0–C=(1)

50=0/2(2)

式中:0为芘的初始负载浓度,μg/g;C为采样时芘的浓度,μg/g;为采样时间,d;为生物降解速率常数,d-1;50为芘的半衰期,d.

采用单因素方差分析(ANOVA)比较空白组(CK)和芘处理组(PYR)之间的差异.所有图使用Origin 2021和R进行绘制.

2 结果与讨论

2.1 芘在土壤中的降解特征

与灭菌土壤相比,在非灭菌土壤环境中,微生物降解是PAHs最主要的衰减模式之一.PAHs在土壤中的降解行为与其结构及理化性质等方面息息相关.本研究选择在同一受试土壤中负载浓度相同且降解条件一致的菲(12.21mg/kg),芘(12.09mg/kg)和苯并[a]芘(8.11mg/kg)作为研究对象,3种PAHs的好氧生物降解如表1所示.菲和苯并[a]芘的好氧微生物降解均符合二级动力学,降解速率随着培养时间的延长而降低,而芘的降解符合零级反应动力学(拟合为:C=0-0.171,2=0.985),降解速率与采样时间无关,其降解半衰期和60d内的降解率均介于菲和苯并[a]芘之间.

表1 菲,芘和苯并[a]芘的物理性质及在土壤中的微生物降解特征

有研究发现,PAHs的分子量与半衰期和疏水性呈正相关,与降解效率和生物可利用性呈负相关[37-39].因此,降解半衰期和疏水性介于菲和苯并[a]芘的芘,其降解效率和生物可利用性也介于两者之间.因此,芘对土壤的吸附作用介于菲和苯并[a]芘之间,导致其生物可利用性介于两者之间.以上结果表明,结构越复杂,疏水性越强的PAHs在土壤中停留时间越长,降解效率和生物可利用性越低.

2.2 芘对土壤酶活性的影响

通过分析脱氢酶和脲酶活性的变化来揭示芘对微生物氧化还原活性及无机氮供应能力的影响.如图1(a)所示,芘处理组中土壤脱氢酶活性在第0,30和60d均显著高于空白组(0.05),促进率为25.4%～49.3%.说明芘在降解第0d和降解后期(第30,60d)刺激了土壤脱氢酶活性.已有报道发现,由于芘为土壤降解菌提供了代谢基质的原因,50～ 200mg/kg的芘不仅可以增加土壤微生物代谢活力,还能刺激土壤脱氢酶活性[41-42].此外,脱氢酶可以参与PAHs的氧化,与PAHs降解速率呈显著正相关[43-44].这说明脱氢酶活性的增强有可能是由于参与了芘的降解氧化过程.脲酶作为尿素水解的关键酶,是土壤氨氮的来源之一[45].如图1(b)所示,与空白组相比,芘处理组仅在第1d显著促进了脲酶活性,促进率为21.4%,说明芘在降解第1d对脲酶活性有一定的刺激作用,但在后期脲酶活性可能适应了芘的胁迫.

图1 空白组和芘处理组中脱氢酶和脲酶活性的变化

图中不同小写字母代表差异显著(<0.05),下同

2.3 芘对土壤氮转化细菌群落组成及功能的影响

微生物群落是执行固氮,氨氧化等土壤氮转化功能的主力军,分析微生物群落结构一定程度上可反映土壤潜在功能的变化,并预测外部压力下的生态系统稳定性.芘对土壤氮转化相关细菌群落组成影响如图2(a)所示,按细菌功能来分,与硝化过程相关的和固氮细菌是所有土壤样品中最丰富的细菌属,占总丰度的74.3%～ 94.4%.与空白组相比,处理组中的相对丰度由第0d的68.7%增长到第3d的89.6%,随后在第60d又回降至61.0%.属于氨氧化古菌,参与土壤中的好氧氨氧化和硝化过程[46].有研究表明,氨氧化微生物对碳氢化合物的污染很敏感[47],且在有机质含量较低的碱性土壤中,是氨氧化过程的主要参与者,具有明显的功能优势[48].因此相对丰度的变化潜在的表明了芘在最初促进了好氧氨氧化和硝化过程,在后期则表现为抑制作用.固氮细菌,[49]和[49]在芘处理初期相对丰度显著降低,在第3d降低至3.5%,随着芘的降解,它们的总相对丰度在第60d又增加至21.0%.其中,属于寡营养型细菌,更倾向于生活在营养物质或有机碳有限的环境中[50].芘的添加可能提供了一定数量的有机碳,导致芘处理过程中的相对丰度呈现出先降低后增加的趋势.此外,[51]和[49]分别还具有尿素分解和硝酸盐还原功能.对于参与尿素分解过程的[51]和硝酸盐还原过程的[52],与空白组相比,其相对丰度在芘胁迫下的前3d由2.3%～2.7%降低至0.2%～0.6%,随后在第60d又升高至1.5%～2.4%.这说明,芘在前期还有可能抑制土壤中的尿素分解和硝酸盐还原过程,但在降解后期又对其表现为促进作用.综上,芘在降解过程中对氨氧化古菌的生长表现为先促进后抑制的作用,而对于固氮细菌,和以及尿素分解细菌和硝酸盐还原细菌则作用相反.

图2 空白组和芘处理组中土壤氮转化细菌属水平上的组成(a)及氮转化相关功能预测结果(b)

根据上述土壤中氮转化微生物群落结构的变化,我们进一步分析了芘对氮转化功能强度的影响.如图2(b)所示,土壤样品中硝化功能(Nitrification)和好氧氨氧化功能(Aerobic_ammonia_oxidation)的细菌丰度最高,占总丰度的72.9%～96.1%.与空白组相比,芘在前3d促进了氨氧化功能和硝化功能的细菌增长,促进率分别为10.5%和10.3%.推测可能是土壤中富集了许多能代谢氨的氨氧化菌,其群落结构的变化导致土壤中相关功能的变化.此外,固氮(Nitrogen_fixation),硝酸盐还原(Nitrate_reduction)和尿素分解(Ureolysis)的功能强度在芘培养的第0d由3.6%～5.8%降低至第3d的0.9%～1.1%,但在第60d又增加至4.3%～7.3%.结合之前的讨论,芘对土壤中氮转化细菌的影响可能会直接导致对应功能强度的变化,最终干扰土壤中的氮转化过程.

2.4 芘对土壤氮转化相关酶活性的影响

为了揭示芘对土壤中氮转化相关酶活性的影响,进一步定量分析了土壤氨氧化和反硝化过程中关键酶活性的变化.如图3(a)所示,在所有采样时间点,芘处理组中AMO活性均显著低于空白组(0.05).表明芘在处理过程中对AMO活性呈现抑制作用,且抑制作用在第7d最大,为56.1%.如图3(b)所示,与空白组相比,芘在第7,30,60d对NAR活性呈显著的抑制作用(0.05).对于NIR活性,芘在第0,1,15,60d对其表现出抑制作用,抑制率在14.0%～38.0%.以上结果表明,与2.3的讨论结果不同,芘在降解前3d增强了土壤氨氧化功能强度,但在整个降解过程中显著抑制了AMO活性,推测可能是由于其他氨氧化菌的减少.对于反硝化过程中的NAR和NIR活性,芘在培养第60d对其均表现为明显的抑制作用.

2.5 芘对土壤氮转化相关基因的影响

为了明确芘对土壤氮转化相关基因的影响,进一步分析了氮转化过程编码基因丰度随时间的变化趋势.如图4(a)所示,空白组和处理组中固氮基因H的丰度第0, 1, 3d有所上升,但到第7d开始下降,且芘对其抑制作用达到最大值,为27.8%,随后抑制作用减弱.与2.3结论不同,土壤中的固氮功能强度在芘胁迫下的前3d呈现出降低的趋势,并未增加.编码基因为AOAA和AOBA分别由氨氧化古菌和氨氧化细菌进行催化.如图4(b),芘仅在第15, 30, 60d显著降低了AOAA基因的丰度,抑制率为47.5%～55.7%.与前面的讨论结果不同,氨氧化古菌的丰度在芘培养第15d和第30d无明显变化.此外,芘在第15d和第60d显著降低了AOBA的丰度,抑制率分别为44.5%和33.8%.说明氨氧化细菌在芘污染环境下也受到显著的抑制作用.如图4(d)所示,芘处理组中编码NAR的G基因丰度仅在第3d显著低于空白组(0.05),在其他时间点与空白组无显著差异.另外,处理组中S基因的丰度在第7d显著低于空白组.这表明芘对这两个基因的丰度影响比较小.NAR可分为膜结合和周质NAR,分别由G和编码[53].本研究观察到芘处理组中NAR活性显著降低,而G基因丰度变化较小,推测可能是芘抑制了G和基因的表达量,这需要进一步探究.NIR同样也有两种类型:可溶性含铜酶和Cu型NIR,分别由S和K编码[53].上述的定量分析结果可知,芘在培养初期(第0, 1, 7d)和后期(第15, 30, 60d)分别显著抑制了固氮基因和氨氧化基因的丰度,但对于反硝化基因G和S,芘对其抑制作用分别仅在第3d和第7d表现明显.另外,本研究发现基因丰度的变化趋势和功能活性不一致,推测可能是由于芘还影响了相关基因的表达量,未来还需要进行深入探究.

图4 空白组和芘处理组土壤微生物氮转化编码基因丰度的变化

3 结论

3.1 芘(12.09mg/kg)在土壤中的好氧微生物降解符合零级反应动力学,降解半衰期为37d.因受其结构和理化性质的影响,其降解率,降解半衰期以及生物可利用性均介于菲和苯并[a]芘之间.

3.2 土壤酶活性分析结果显示,芘仅在降解第1d刺激了脲酶活性,而在降解第0d和降解后期(第30, 60d)显著刺激了土壤脱氢酶活性.这说明环境浓度的芘整体上促进了微生物的代谢活力.

3.3 分析土壤中氮转化相关微生物的结构变化可知,由于氨氧化古菌相对丰度的变化,芘在培养前期(第0, 1, 3d)促进了好氧氨氧化和硝化功能,在第60d则表现为抑制作用,而对于具有固氮功能的,以及,尿素分解细菌和硝酸盐还原细菌则作用相反.因此,芘对土壤中氮转化相关微生物群落结构的影响可能会导致对应功能的变化.

3.4 从功能基因方面定量分析芘对土壤氮转化过程的影响,结果显示:在芘处理前3d固氮基因H的丰度有所增加,但第7d芘对其表现出的抑制作用达到最大,随后抑制作用减弱.根据土壤氨氧化和反硝化过程中关键酶活性及编码基因的变化,芘在降解过程中显著抑制了氨氧化过程,但对于氨氧化基因(AOAA和AOBA)仅在降解后期(第15, 30, 60d)起到抑制作用.此外,芘对反硝化过程也主要在降解第60d表现为抑制作用,但对其编码的G和S的基因丰度则作用不明显.

3.5 氮转化基因定量分析结果与相关微生物的群落结构和功能预测结果不同.因此,为了明晰污染物对元素转化过程的影响特征,不仅需要定量分析相关酶活性以及基因,未来还需要进一步定量分析基因的RNA表达量,才能准确评估微生物群落结构与功能活性的变化.

[1] 孙 娇,张作涛,郭海礁,等.多环芳烃厌氧生物降解研究进展[J]. 微生物学报, 2020,60(12):2844-2861. Sun J, Zhang Z T, Guo H J, et al.Progresses in anaerobic microbial degradation of polycyclic aromatic hydrocarbons [J].Acta Microbiologica Sinica, 2020,60(12):2844-2861.

[2] Flowers L, Rieth S H, Cogliano V J. et al. Health assessment of polycyclic aromatic hydrocarbon mixtures: Current practices and future directions [J]. Polycyclic Aromatic Compounds, 2002,22(3/4): 811-821.

[3] 曾 军,吴宇澄,林先贵.多环芳烃污染土壤微生物修复研究进展[J]. 微生物学报, 2020,60(12):2804-2815. Zeng J, Wu Y C, Lin X G.Advances in microbial remediation of soils polluted by polycyclic aromatic hydrocarbons [J]. Acta Microbiologica Sinica, 2020,60(12):2804-2815.

[4] Maisto G, Nicola F D, Lovieno P, et al. PAHs and trace elements in volcanic urban and natural soils [J]. Geoderma, 2006,136(1):20-27.

[5] Liu S D, Xia X H, Yang L Y, et al. Polycyclic aromatic hydrocarbons in urban soils of different land uses in Beijing, China: Distribution, sources and their correlation with the city's urbanization history [J]. Journal of Hazardous Materials, 2010,177(1):1085-1092.

[6] Huang Z Y, Liu Y, Dai H, et al. Spatial distribution and source apportionment of polycyclic aromatic hydrocarbons in typical oasis soil of north-western China and the bacterial community response [J]. Environmental Research, 2022,204:112401.

[7] 尚庆彬,段永红,程 荣.中国农业土壤多环芳烃污染现状及来源研究[J]. 山东农业科学, 2019,51(3):62-67. Shang Q B, Duan Y H, Cheng R.Pollution status and sources of polycyclic aromatic hydrocarbons in agricultural soils in China [J]. Shandong Academy of Agricultural Sciences, 2019,51(3):62-67.

[8] Wang D, Zhu S L, Wang L J. et al. Distribution, origins and hazardous effects of polycyclic aromatic hydrocarbons in topsoil surrounding oil fields: A case study on the Loess Plateau, China [J]. International Journal of Environmental Research and Public Health, 2020,17(4): 1390.

[9] Ouyang Z, Gao L, Yang C. Distribution, sources and influence factors of polycyclic aromatic hydrocarbon at different depths of the soil and sediments of two typical coal mining subsidence areas in Huainan, China [J]. Ecotoxicology and Environmental Safety, 2018,163: 255-265.

[10] Enuneku A, Ogbeide O, Okpara B, et al. Ingestion and dermal cancer risk via exposure to polycyclic aromatic hydrocarbon-contaminated soils in an oil-producing community, Niger Delta, Nigeria [J]. Environmental Toxicology and Chemistry, 2021,40(1):261-271.

[11] Sun Z, Zhu Y, Zhuo S. J,et al.Occurrence of nitro- and oxy-PAHs in agricultural soils in eastern China and excess lifetime cancer risks from human exposure through soil ingestion [J]. Environment International, 2017,108:261-270.

[12] Wang N, Li H B, Long J L,et al. Contamination, source, and input route of polycyclic aromatic hydrocarbons in historic wastewater- irrigated agricultural soils [J]. Journal of Environmental Monitoring, 2012,14(12):3076-3085.

[13] Zhang X X, Zhu C D, Wang F,et al.Pollution characteristics and risk assessment of polycyclic aromatic hydrocarbons in agricultural soils of different land use types in Nanjing Suburbs [J]. Environmental Earth Sciences, 2023,44(2):944-953.

[14] Fuhrman J A. Microbial community structure and its functional implications [J]. Nature, 2009,459(7244):193-199.

[15] Vlek P L G, Fillery I R P, Burford J R. Accession, transformation, and loss of nitrogen in soils of the arid region [J]. Plant and Soil, 1981, 58(1-3):133-175.

[16] Sverdrup L E, Ekelund F, KroghP H, et al. Soil microbial toxicity of eight polycyclic aromatic compounds: Effects on nitrification, the genetic diversity of bacteria, and the total number of protozoans [J]. Environmental Toxicology and Chemistry, 2002,21(8):1644-1650.

[17] Yi M L, Zhang L L,Qin C L, et al. Temporal changes of microbial community structure and nitrogen cycling processes during the aerobic degradation of phenanthrene [J]. Chemosphere, 2022,286(2):131709.

[18] Yi M L, Zhang L L, Li Y, et al. Structural, metabolic, and functional characteristics of soil microbial communities in response to benzo a pyrene stress [J]. Journal of Hazardous Materials, 2022,431:128632.

[19] Chen Y Y, Zhu L Z, Zhou R B.Characterization and distribution of polycyclic aromatic hydrocarbon in surface water and sediment from Qiantang River, China [J]. Journal of Hazardous Materials, 2007, 141(1):148-155.

[20] Crampon, M, Bureau F, Akpa-Vinceslas M, et al. Correlations between PAH bioavailability, degrading bacteria, and soil characteristics during PAH biodegradation in five diffusely contaminated dissimilar soils. Environmental Science and Pollution Research, 2014,21(13):8133-8145.

[21] Guo G, Tian F, Ding K Q, et al. Effect of a bacterial consortium on the degradation of polycyclic aromatic hydrocarbons and bacterial community composition in Chinese soils [J]. International Biodeterioration & Biodegradation, 2017,123:56-62.

[22] Belkin S, Stieber M, Tiehm A, et al.Toxicity and genotoxicity enhancement during polycyclic aromatic hydrocarbons' biodegradation [J].Environmental Toxicology and Water Quality, 1994,9(4):303-309.

[23] Kanaly R A, Harayama S. Biodegradation of high-molecular-weight polycyclic aromatic hydrocarbons by bacteria [J]. Journal of Bacteriology,2000,182(8):2059-2067.

[24] Sun F L, Wang Y S, Sun C C, et al. Effects of three different PAHs on nitrogen-fixing bacterial diversity in mangrove sediment. Ecotoxicology, 2012,21(6):1651-1660.

[25] 刘 杰.土壤多环芳烃(PAHs)污染对氨氧化菌及微生物群落结构的影响研究[D]. 北京:北京林业大学, 2011:2-22. Liu J. The influence on ammonia oxidizing microorganisms and microbial community structure under soil PAHs pollution [D]. Beijing:Beijing Forestry University, 2011:2-22.

[26] Ige I D, Olutona G O, Ajaelu C J. Insight into the metropolitan levels, spatial distribution and health risks of polycyclic aromatic hydrocarbons in roadside soil of Ibadan, Nigeria [J]. Environmental Earth Sciences, 2021,80(20):687.

[27] Raudonytė-Svirbutavičienė E R, Stakėnienė R, Jokšas K, et al. Distribution of polycyclic aromatic hydrocarbons and heavy metals in soil following a large tire fire incident: A case study [J]. Chemosphere, 2022,286(1):131556.

[28] Sun N, Liu Q, Wang J H, et al. Probing the biological toxicity of pyrene to the earthworm Eisenia fetida and the toxicity pathways of oxidative damage: A systematic study at the animal and molecular levels [J]. Environmental Pollution, 2021,289:117936.

[29] Wang J, Luo Z J, Song Y Y, et al. Remediation of phenanthrene contaminated soil by g-C3N4/Fe3O4composites and its phytotoxicity evaluation [J]. Chemosphere, 2019,221:554-562.

[30] Sinsabaugh R L, Hill B H, Follstad Shah J J. Ecoenzymatic stoichiometry of microbial organic nutrient acquisition in soil and sediment [J]. Nature, 2009,462(7274):795-798.

[31] Li X N, Qu C S, Bian Y R, et al. New insights into the responses of soil microorganisms to polycyclic aromatic hydrocarbon stress by combining enzyme activity and sequencing analysis with metabolomics [J]. Environmental Pollution, 2019,255(Pt 2):113312.

[32] 邢 奕,王志强,李益飞,等.不同粒度CuO及与乙基黄原酸钾复合污染对土壤脲酶和微生物多样性的影响 [J]. 中国环境科学, 2017, 37(4):1466-1473. Xing Y, Wang Z Q, Li Y F, et al. Effects of different sizes of CuO and potassium ethyl potassium compound pollution on soil urease and microbial diversity [J]. China Environmental Science, 2017,37(4): 1466-1473.

[33] Kaczynska G, Borowik A, Wyszkowska J. Soil dehydrogenases as an indicator of contamination of the environment with petroleum products [J]. Water Air and Soil Pollution, 2015,226(11):372.

[34] Junier, P, Molina, V, Dorador, C, et al. Phylogenetic and functional marker genes to study ammonia-oxidizing microorganisms (AOM) in the environment [J]. Appl Microbiol Biotechnol, 2010,85(3):425-440.

[35] Yang X, He Q, Guo F C, et al. Nanoplastics disturb nitrogen removal in constructed wetlands: Responses of microbes and macrophytes [J]. Environmental Science & Technology, 2020,54(21):14007-14016.

[36] Louca S, Parfrey L W, Doebeli M J S, Decoupling function and taxonomy in the global ocean microbiome [J]. Science, 2016,353: 1272-1277.

[37] Shi Z, Tao S, Pan B, et al. Contamination of rivers in Tianjin, China by polycyclic aromatic hydrocarbons [J]. Environmental pollution, 2005, 134:97-111.

[38] Roslund M I, Grönroos M, Rantalainen A L, et al. Half-lives of PAHs and temporal microbiota changes in commonly used urban landscaping materials [J].PeerJ, 2018,6(1):4508.

[39] Park K S, Sims R C, Dupont R R. Transformation of PAHs in soil systems [J]. Environmental Engineering, 1990,116:632-640.

[40] Zhang L L, Yi M L, Lu P L. Effects of pyrene on the structure and metabolic function of soil microbial communities [J]. Environmental Pollution, 2022,305:119301.

[41] Wang C, Luo Y, Tan H, et al. Responsiveness change of biochemistry and micro-ecology in alkaline soil under PAHs contamination with or without heavy metal interaction [J]. Environmental Pollution, 2020, 266(3):115296.

[42] Lu M, Xu K, Chen J. Effect of pyrene and cadmium on microbial activity and community structure in soil [J]. Chemosphere, 2013, 91(4):491-497.

[43] Ma, L, Deng F C, Yang C, et al. Bioremediation of PAH-contaminated farmland: field experiment [J]. Environmental Science and Pollution Research, 2018,25(1):64-72.

[44] Kaimi E, Mukaidania T, Miyoshia S J, et al. Ryegrass enhancement of biodegradation in diesel-contaminated soil [J]. Environmental and Experimental Botany, 2006,55(1/2):110-119.

[45] Ibrahim M M, Tong C X, Hu K, et al. Biochar-fertilizer interaction modifies N-sorption, enzyme activities and microbial functional abundance regulating nitrogen retention in rhizosphere soil [J]. Science of the Total Environment, 2020,739:140065.

[46] Stieglmeier M, Klingl A, Alves R J E, et al. Nitrososphaera viennensis gen. nov., sp nov., an aerobic and mesophilic, ammonia-oxidizing archaeon from soil and a member of the archaeal phylum Thaumarchaeota [J]. International Journal of Systematic and Evolutionary Microbiology, 2014,64:2738-2752.

[47] Yang Y C, Herbold C W, Jung M Y, et al. Survival strategies of ammonia-oxidizing archaea (AOA) in a full-scale WWTP treating mixed landfill leachate containing copper ions and operating at low-intensity of aeration [J]. Water Research, 2021,191:116798.

[48] Sim J X F,Doolette C L,Vasileiadis S, et al. Pesticide effects on nitrogen cycle related microbial functions and community composition [J]. Science of The Total Environment, 2022,807:150734.

[49] Claudine F, Kristina L, Claudine E. Nitrogen-fixing bacteria associated with leguminous and non-leguminous plants [J]. Plant and Soil, 2009,321(1):35-59.

[50] Ohta H, Hattori T.gen. nov., sp. nov., a nitrogen-fixing oligotrophic bacterium [J]. Antonie Van Leeuwenhoek Journal of Microbiology, 1983,49(4/5):429-446.

[51] Brenner D J, Krieg N R, Staley J T, et al.The, Part B: The. Bergey's Manual of Systematic Bacteriology [M]. New York, Springer, 2005,Vol.2.

[52] Brenner D J, Krieg N R, Staley J T, et al. The Proteobacteria, Part B: The Gammaproteobacteria. Bergey's Manual of Systematic Bacteriology [M]. New York, Springer, 2011,Vol.4.

[53] 郭丽芸,时 飞,杨柳燕.反硝化菌功能基因及其分子生态学研究进展[J]. 微生物学通报, 2011,38(4):583-590. Guo L Y, Shi F, Yang L Y. Advances in functional genes and molecular ecology in denitrifiers[J], Microbiology China, 2011,38(4):583-590.

Characterization of pyrene's impact on the soil functional microorganisms for nitrogen transformation.

HU Qin1, ZHANG Li-lan1,2*, YI Mei-ling1, YANG Rui1

(1.Key Laboratory of Three Gorges Reservoir Region’s Eco-environment, Ministry of Education, Chongqing University, Chongqing 400044, China；2.State Key Laboratory of Coal Mine Disaster Dynamics and Control, Chongqing University, Chongqing 400044, China)., 2023,43(10)：5574～5582

An aerobic degradation microenvironment was constructed to analyze the effects of pyrene (12.09mg/kg) at ambient concentration on soil enzyme activities, the whole process of nitrogen transformation and related functional microorganisms. The results showed that pyrene only significantly increased urease activity on the first day of degradation, but promoted the dehydrogenase activity at both the early and late phases of degradation. The analysis of the bacterial community structure revealed that the variation of the relative abundance of ammonia-oxidizing bacteria () promoted pyrene-mediated aerobic ammonia oxidation and nitrification in the early stages of treatment and inhibited that in the late stages. In contrast, the effects of nitrogen-fixing bacteria (,and), urea-degrading bacteria (), and nitrate-reducing bacteria () were opposite. The quantitative analysis of functional genes showed that, despite pyrene's repressive effect on the nitrogen-fixing geneH at the start of the culture, the abundance ofH showed an increasing trend, which was not consistent to the anticipated changes in microbial community structure and associated functions. Compared with changes in key enzyme activities and genes encoding the processes of ammonia oxidation and denitrification, pyrene did not significantly boost ammonia oxidation in the early stages of incubation, t severely hampered ammonia oxidation and denitrification after 15days, significantly inhibited the ammonia oxidation process. In this study, how pyrene influenced the microbial nitrogen transformation process in soil was reported, fundamental data on understanding the environmental hazard of pyrene were provided.

pyrene；soil enzyme activity；nitrogen cycling bacterial community；nitrogen cycling processes

X172,X171.5

A

1000-6923(2023)10-5574-09

2023-03-01

国家重点研发计划(2019YFC1805500);国家自然科学基金资助项目(42177363)

* 责任作者, 副教授, lilanzhang@cqu.edu.cn

胡 琴(1997-),女,重庆忠县人,重庆大学硕士研究生,主要研究方向为抗性基因分子生物学.zh3357@qq.com.

胡 琴,张利兰,易美玲,等.芘对土壤微生物氮转化功能菌群的影响特征 [J]. 中国环境科学, 2023,43(10):5574-5582.

Hu Q, Zhang L L, Yi M L, et al. Characterization of pyrene's impact on the soil functional microorganisms for nitrogen transformation [J]. China Environmental Science, 2023,43(10):5574-5582.