微生物学通报  2020, Vol. 47 Issue (10): 3342−3354
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随着我国人民生活水平的不断提高以及人均用水量的不断增大，居民对水质的要求也愈发严格。《2018年中国环境状况公报》[1]显示，氮素污染仍然是我国现阶段水环境中的主要污染物之一。传统的生物脱氮技术依赖于硝化作用(nitrification)和反硝化作用(denitrification)的有机结合，但由于硝化和反硝化作用的功能微生物其生态位差别非常大，硝化和反硝化作用通常需要在两个独立的反应器中进行。好氧反硝化(aerobic denitrification)微生物的发现，使得同步硝化反硝化(simultaneous nitrification and denitrification，SND)成为了可能，这不仅大大节省了基建投资费用，反硝化过程产生的碱度也能够补偿硝化作用所消耗的酸度，进一步减少了运行成本。因此，凭借其独特优势，好氧反硝化生物脱氮技术自提出以来，受到国内外环境领域学者的广泛关注。
好氧反硝化的概念是由荷兰学者Robertson和Kuenen教授共同提出的[2]。基于对好氧系统内发生明显异化的总氮损失的疑惑，两位学者成功地从一反硝化脱硫脱氮反应器中筛选分离出一株好氧反硝化菌Thiosphaera pantotropha (后改名为Paracoccus pantotrophus)[3]，经过一系列进一步的实验验证，从而正式提出好氧反硝化的概念[4]。本文总结了近年来好氧反硝化菌的筛选分离情况及好氧反硝化菌脱氮特性方面的研究进展，同时简要介绍好氧反硝化生物脱氮技术原理及实际应用方面的研究现状。最后，对未来好氧反硝化生物脱氮技术的研究方向提出了科学的展望。
1 好氧反硝化菌种的分离现状 1.1 筛菌方法获得好氧反硝化菌种资源的步骤一般可分为富集驯化阶段以及筛选分离阶段。富集驯化阶段一般多采用间歇曝气法[5]，而在筛选分离阶段一般可采用溴麝香草酚(bromothymol blue，BTB)培养基法或其他筛选培养基法。王弘宇等[6]提出了一种高效筛选好氧反硝化菌的方法，即依次经过污泥驯化、细菌分离纯化、初筛[测总氮(total nitrogen，TN)]、复筛(氮元素轨迹跟踪测定法)等阶段，其筛选出的好氧反硝化菌Pseudomonas chloritidismutans X31不仅具有较好的好氧反硝化效果，同时对氧气的耐受性也特别好[7]。高珊珊[8]通过定量投加硝酸盐并控制反应器曝气强度来驯化活性污泥，以提高活性污泥在好氧条件下对硝酸盐的去除率，达到富集好氧反硝化菌的目的，成功筛选出5株快速去除硝酸盐的好氧反硝化细菌，并扩增出与好氧反硝化作用密切相关的周质硝酸盐还原酶编码基因napA。
1.2 菌种分离情况目前报道的已分离好氧反硝化菌种类繁多，如不动杆菌属(Acinetobacter sp.)、气单胞菌属(Aeromonas sp.)、副球菌属(Paracoccus sp.)及假单胞菌属(Pseudomonas sp.)等。另一方面，好氧反硝化菌种来源也不尽相同，如活性污泥、土壤、沉积物、海洋湖泊及各类废水等，这说明好氧反硝化菌在自然界中是广泛存在的。近年来，一系列具有特殊功能的好氧反硝化菌被筛选分离出来，如具有重金属抗性的好氧反硝化菌[9]、耐酸性的好氧反硝化菌[10]、耐盐的好氧反硝化菌[11]、具有聚磷特性的好氧反硝化菌[12]、具有降解难降解有机物的好氧反硝化菌[13]和耐低碳氮比(C/N)的好氧反硝化菌[14]等。近年来分离出的好氧反硝化菌种类及来源如表 1所示。
| 种属 Species |
菌名 Strain |
来源 Source |
年份 Year |
| Agrobacterium sp. | LAD9 | Landfill leachate treatment system | 2011[15] |
| Brevibacterium | yy7 | A2/O wastewater treatment plant | 2011[16] |
| Psychrobacter sp. | S1-1 | Biological aerated filter | 2011[17] |
| Pseudomonas stutzeri | YZN-001 | Piggery wastewater treatment system | 2011[18] |
| Bacillus methylotrophicus | L7 | Wastewater sample | 2012[19] |
| Rhodococcus sp. | CPZ24 | Swine wastewater | 2012[20] |
| Halomonas campisalis | ha3 | Saline-alkali lake | 2013[21] |
| Paracoccus versutus | LYM | Seabed sludge | 2013[22] |
| Chryseobacterium sp. | R31 | Slaughterhouse wastewater | 2014[23] |
| Aeromonas sp. | HN-02 | CASS reactor | 2014[24] |
| Pseudomonas stutzeri | PCN-1 | Biological aerated filter | 2014[25] |
| Alcaligenes faecalis | C16 | Aeration tank | 2015[26] |
| Klebsiella pneumonia | EGD-HP19-C | Industrial wastewater | 2015[27] |
| Diaphorobacter sp. | PD-7 | Coking-plant wastewater ponds | 2015[28] |
| Vibrio diabolicus | SF16 | Marine sediment | 2015[29] |
| Pseudomonas aeruginosa | PCN-2 | Landfill leachate treating reactor | 2015[30] |
| Pseudomonas stutzeri | C3 | Wastewater treatment plant | 2015[31] |
| Pseudomonas tolaasii | Y-11 | Long-term flooded paddy soil | 2016[32] |
| Cupriavidus sp. | S1 | Coking wastewater | 2016[33] |
| Pseudomonas brassicacearum | LZ-4 | Petrochemical wastewater | 2016[34] |
| Diaphorobacter polyhydroxybutyrativorans | SL-205 | Denitrification reactor | 2017[35] |
| Enterobacter cloacae | HW-15 | Phosphorus-rich river | 2017[36] |
| Acinetobacter sp. | H36 | Sediment | 2017[37] |
| Raoultella sp. | R11 | Eutrophic lake | 2017[38] |
| Pseudomonas putida | Y-9 | Long-term flooded paddy soil | 2017[39] |
| Acinetobacter sp. | JR1 | Pharmaceutical raw water | 2019[10] |
| Pseudomonas putida | NP5 | Activated sludge | 2019[40] |
| Acinetobacter johnsonii | WGX-9 | Sediment of a drinkingwater reservoir | 2019[41] |
| Paracoccus sp. | YF1 | Activated sludge | 2019[42] |
分析表 1能够发现，近年来所筛选分离出的好氧反硝化菌主要集中在变形菌门(Proteobacteria)。不同种属细菌的生长及脱氮特性一般不同，即使同种类型的细菌，在代谢过程中往往也存在差异。
2 环境因子对好氧反硝化影响的研究进展溶解氧(dissolved oxygen，DO)、C/N、温度、碳源种类、pH值等因素通常被认为是影响好氧反硝化过程的重要环境因素，但是不同种类的菌种、反应器类型及环境条件使这些因素的影响效果不尽相同。近年来的研究成果表明，DO、C/N及温度是其中的关键性限制因素。
2.1 DO一方面，由于在好氧反硝化的过程中硝酸盐和氧均能作为电子受体，二者必然竞争电子。另一方面，从热力学的角度来讲，相比于硝酸盐，氧作为电子受体时产生的能量更多，更易被微生物利用。因此，DO必然会在一定程度上影响好氧反硝化过程，影响程度可能根据微生物种类的不同而存在差异。
根据多年来报道的好氧反硝化菌来看，DO对其好氧反硝化效能的影响可分为三类：(1)有些好氧反硝化菌的好氧反硝化效能随DO的升高而降低，当DO超过某一值时，DO浓度的升高便不再影响好氧反硝化菌的反硝化效能[43]，该DO的浓度值称为其DO阈值；(2)大部分好氧反硝化菌在某一DO范围内具有较高的好氧反硝化效能，低于或高于该范围，好氧反硝化效能均会下降[44]，该DO范围随微生物种类不同可发生变化；(3)还有一少部分好氧反硝化菌能够耐受较高浓度的DO，如P. stutzeri YZN-001[18]和P. stutzeri X31[45]，即在较高DO时仍然具有较高的好氧反硝化效能。
然而，关于DO影响好氧反硝化菌脱氮效能的机制方面的报道表明，DO是通过影响反硝化酶系的表达，从而影响好氧反硝化脱氮效能及反硝化终产物类型的[25]。反硝化酶系对于DO的耐受程度不同，如亚硝酸盐还原酶及一氧化氮还原酶对氧气较为敏感，因此当DO过高时易造成亚硝酸盐或温室气体N2O的积累[46]。
2.2 C/N目前发现的绝大多数好氧反硝化菌均是异养菌，即利用有机物作为碳源和能源。因此有机碳的多少势必会影响好氧反硝化作用。研究表明，在适当的范围内，作为能源的碳源浓度越高，好氧反硝化速率越快[47]。大部分好氧反硝化菌的最佳C/N为4-5[48]，有的菌可能会相对较高，达到9-10[44]。相比于传统的缺氧反硝化所需C/N，好氧反硝化需要的C/N更高。当然，目前也筛选分离出一些耐低C/N比的好氧反硝化菌(群)[14]，该类菌一般多从养殖水体、水库水体等贫营养条件下筛选得到。
2.3 温度温度不仅能够通过影响酶的活性而影响微生物的代谢，而且还能够影响反应的活化能。因此，温度对于好氧反硝化菌也具有显著的影响。大部分好氧反硝化菌的最适温度为25-30 ℃[47]。低温或者高温会严重抑制好氧反硝化菌的生长及活性。然而目前筛选出了多种不同种属的耐低温[49-53]或耐高温[54-56]好氧反硝化菌，为好氧反硝化生物脱氮技术的进一步实际应用提供了可能。
3 好氧反硝化机理的研究现状好氧反硝化是指在好氧条件下能够进行的反硝化过程，这与传统观念认为反硝化过程只能发生在严格厌氧或缺氧条件下的观念相悖。截至目前，人们对于好氧反硝化的反应机理仍然存在不同的观点，对其机理的探讨还未达到令人满意的程度，尤其是在好氧反硝化的生理学及生态学意义上的讨论仍然存在一定的争议[57]。但综合分析近年来相关的研究成果和理论，可以从微环境和微生物学的角度对其加以解释[58]。
3.1 微环境理论微环境理论主要侧重从物理学角度对好氧反硝化作用机理进行解释[59]，该理论认为在微生物絮体内，由于氧的扩散受到限制，在微生物絮体内产生DO梯度，微生物絮体外表面DO较高，以好氧异养菌、好氧硝化菌为主；随着氧向絮体内部传递，传递阻力增大的同时，有机物氧化、硝化作用也消耗大量氧，导致絮体内部产生缺氧区，反硝化菌占优势。正是由于微生物絮体内缺氧微环境存在，从而导致微环境中反硝化作用的发生。将曝气池内DO控制在较低水平，将可能提高缺氧或厌氧微环境所占比例，从而促进反硝化作用。实际上，由于微生物种群结构、基质分布代谢活动和生物化学反应的不均匀性，以及物质传递的变化等因素相互作用，在微生物絮体和生物膜内部会存在多种多样的微环境。
微环境理论从微观的角度解释了宏观发生的“好氧反硝化”作用，但本质上仍然是建立在反硝化过程发生于缺氧条件下的基础上，结合了客观存在的现象而加以解释，严格意义上来说并不是真正意义上的好氧反硝化，该理论同样用来解释同步硝化反硝化现象。
3.2 微生物学理论微生物学理论主要侧重从微生物本身的性质出发对好氧反硝化过程予以解释。该理论认为，好氧反硝化过程是在好氧反硝化菌的作用下完成的，好氧反硝化菌能够以氧气和硝酸盐同时作为电子受体，即Robertson等提出的“协同呼吸(co-respiration)”理论[2]，认为在电子传递给氧气的过程中，电子传递链中的某一环节存在“瓶颈”，即限速步骤，导致过量的电子能够传递给反硝化酶系，从而发生好氧反硝化作用。好氧反硝化菌的电子传递链示意图如图 1所示。
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| 图 1 好氧反硝化菌电子传递链示意图 Figure 1 Schematic diagram of electron transfer chain of aerobic denitrification |
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Chen等[60]认为微生物的呼吸链高度分支化，这个复杂的呼吸链允许微生物构建一个最具能效的途径来适应环境条件。当微生物对能量的需求较少时，由于氧气的电子亲和力高于硝酸盐，微生物会更倾向于选择氧气作为电子受体；当微生物对能量的需求较大时，硝酸盐作为电子受体的竞争力提升，电子传递给硝酸盐的比例会增大。随着周质硝酸盐还原酶(periplasmic nitrate reductase，Nap)在好氧条件下依然能表达的性质的发现，进一步证明了好氧反硝化作用存在发生的可能性。
综上，微环境理论着眼于实际混菌(如活性污泥、生物膜)条件下氮的去除机制，是从实际反应器角度出发提出的好氧反硝化的机理。微生物学理论是从微生物本身特性的角度提出的好氧反硝化机理。近年来，关于好氧反硝化菌代谢机理方面的研究成果较少[61]，尤其是在电子传递机制方面的研究有待进一步加强。
4 好氧反硝化分子生物学及酶学的研究进展反硝化反应是由反硝化酶系(硝酸盐还原酶、亚硝酸盐还原酶、一氧化氮还原酶及氧化亚氮还原酶)催化的由硝酸盐逐步还原至N2的过程。因此，随着技术手段的不断发展，好氧反硝化研究过程中涉及到的分子生物学以及相关酶学的研究也是近年来研究的热点之一。
4.1 硝酸盐还原酶原核生物硝酸盐还原酶根据其在细胞中的位置及生理功能的不同，可分为膜质硝酸盐还原酶(respiratory nitrate reductase，Nar)、Nap及同化硝酸盐还原酶(assimilatory nitrate reductase，Nas)[62]。Nar与Nap在同一菌体中可同时表达，菌体的好氧生长和缺氧生长直接影响了这两种酶的活性。在缺氧条件下，Nar表达占主导地位，而且仅在缺氧条件下才能发挥作用，在好氧条件下Nap表达占主导地位，并且在好氧和缺氧条件下都能发挥作用[63]。因此，有学者提出，好氧反硝化的第一步反应过程正是由Nap催化调控的[64]。因此，nap基因一直被认为是好氧反硝化作用的关键功能基因。
目前获得的Nap的晶体结构多为NapAB二聚体，其中NapA为催化亚基，负责催化硝酸盐还原为亚硝酸盐的反应，NapB为电子传递亚基，负责将电子传递给NapA[65-66]。最近的研究表明，通过调节NapA亚基中铁硫聚簇([4Fe-4S])的氧化还原电位能够调控Nap的活性[67]。
由于部分好氧反硝化菌具有同时去除化学需氧量(chemical oxygen demand，COD)和NH4+-N的能力，因此被推测具有异养硝化效能。然而，Sun等研究表明，好氧反硝化菌P. stutzeri T13仅利用NH4+-N进行合成代谢，该菌株并不具有异养硝化效能，该菌株在NH4+-N和NO3--N共存时，优先利用NH4+-N为氮源进行合成代谢作用；当NH4+-N耗尽后，微生物能够以NO3--N为氮源进行合成代谢作用[68]。这是由于NH4+-N抑制了Nas的表达。好氧反硝化菌的同化硝酸盐还原作用一直不被重视，是由于有些学者认为微生物的同化量不会太高。但近年有研究报道，一些好氧反硝化菌，尤其是来源于海洋中的微生物，能够达到几乎100%的硝酸盐同化量，对好氧条件下硝酸盐的去除有很大的潜能[69]。
4.2 亚硝酸盐还原酶根据辅基的不同，亚硝酸盐还原酶可分为铜型亚硝酸盐还原酶(Cu-type nitrite reductase，Cu-Nir)以及细胞色素型亚硝酸盐还原酶(cd1-type nitrite reductase，cd1-Nir)[70]。过去的研究表明，这两种酶的编码基因一般不能同时存在于一种微生物中，但是最近的研究发现，这两种酶的编码基因能够同时存在于同一微生物中，但是不能同时表达[15]。Sánchez等利用基因敲除技术证明了在缺氧条件下这两种酶的功能是冗余的[71]，但没有进一步探究在好氧条件下这两种酶的编码基因的表达情况。亚硝酸盐还原酶对氧气较为敏感，因此在好氧反硝化过程中，亚硝酸盐积累一直是一个问题[72]。因此，若能利用基因编译等方法，调节胞内电子传递路线，提高胞内电子传递效率，对好氧反硝化效能的提高具有重要意义，然而该方面的报道仍然比较有限。
4.3 一氧化氮还原酶一氧化氮(nitric oxide，NO)还原酶催化NO还原为氧化亚氮(nitrous oxide，N2O)的反应。由于NO还原酶活性较强，所以NO一般不易积累在系统中。Gui等[73]探究了磺胺甲恶唑(sulfamethoxazole，SMX)对好氧反硝化菌(P. stutzeri PCN-1)脱氮特性的影响，结果表明尽管SMX浓度的改变对PCN-1的cnorB基因的表达量有显著影响，但是对NO的实际产量基本没有任何影响，系统中NO-N的含量始终接近于0 mg/L。Zheng等[25]考察了O2浓度对PCN-1脱氮特性的影响，结果表明系统中最大的NO积累量仍低于被还原硝酸盐的0.003%。
4.4 N2O还原酶N2O是一种破坏效应极强的温室气体。近年来，关于N2O的产生以及环境因子对N2O还原酶编码基因表达的影响，逐渐成为研究的热点。好氧反硝化菌PCN-1在不同O2含量下N2O的最大积累量仍然低于被还原硝酸盐的0.33%[25]。Gui等[74]考察了4种重金属(Cd2+、Cu2+、Ni2+、Zn2+)对好氧反硝化菌PCN-1脱氮特性的影响，结果表明重金属离子浓度的升高会明显抑制nosZ基因的表达，从而导致N2O气体的积累。Pan等通过动力学模型探究反硝化过程中的电子竞争问题，结果表明N2O的半饱和常数要远高于其余3种反硝化底物的半饱和常数，这说明N2O还原酶在电子竞争方面的能力比较强[75]。
总结已有的研究发现，关于好氧反硝化过程的分子生物学研究内容主要集中在好氧反硝化代谢相关酶和基因的结构和功能方面，对于酶分子动力学、细胞内电子传递机制及好氧反硝化代谢调控等方面的研究极为有限。
5 分子生态学技术在好氧反硝化研究中的应用随着分子生态学、生物信息学等学科的飞速发展，利用新兴的技术手段来研究好氧反硝化的报道逐渐增多。如康鹏亮等[76]利用高通量测序技术研究湖库沉积物中好氧反硝化菌群的种群结构情况，所筛选出的3组好氧反硝化优势混合菌群种群结构差异显著，3组菌群的优势菌种分别为Bacillus subtilis、P. pantotrophus和P. stutzeri。Zhou等[77]利用高通量测序技术进一步揭示了水库中氮素的损失主要是由好氧反硝化菌的脱氮作用实现的。Liu等[78]利用基因组测序技术分析好氧反硝化菌Achromobacter sp. GAD3和Agrobacterium sp. LAD9的代谢潜能。Jin等[79]利用荧光定量聚合酶链式反应(quantitative polymerase chain reaction，qPCR)技术从mRNA水平探究了一株异养硝化-好氧反硝化菌(Klebsiella sp. KSND)的氮代谢途径，该菌株同时具有3种硝酸盐还原酶，NapA型硝酸盐还原酶主要负责调控硝酸盐的同化及反硝化过程。Hu等[80]利用高通量测序技术探究了碳源及反应器运行方式对好氧反硝化工艺中微生物群落结构变化的影响，与运行方式相比，碳源的改变对好氧反硝化过程和微生物群落的影响更显著。Fu等[81]利用高通量测序技术探究复合发酵液对人工湿地系统中氮素的去除效能及微生物群落结构变化的影响，结果发现大量的异养硝化菌和好氧反硝化菌存在于人工湿地系统的上层和中层。Du等[82]利用高通量测序技术解析一株假单胞菌(Pseudomonas sp.)强化煤质乙二醇工业废水脱氮处理后，系统内微生物群落结构及功能微生物的变化情况，结果表明假单胞菌属的丰度低于10%时，系统的硝酸盐及亚硝酸盐去除效能明显降低。
6 好氧反硝化生物脱氮技术的应用进展好氧反硝化生物脱氮技术应用方面的研究，主要集中在实验室小试阶段，中试及厂试规模的报道较少。研究中多以生物强化的手段为主，即将好氧反硝化菌(群)以菌剂的形式外源投加于反应器中，以期提高反应器的脱氮效能。也有学者通过外部刺激等手段，以提高好氧反硝化菌(群)本身活性为目的，从而提高反应器的脱氮效能[83]。
近年来关于好氧反硝化生物脱氮技术应用方面的报道主要有：左薇[84]将异养硝化-好氧反硝化菌P. stutzeri AD3投加于序批式反应器(sequencing batch reactor，SBR)中，经过15 d的驯化反应器即达到稳定，对氨氮和TN的去除率可分别达到90%和70%。高珊珊[8]采用活性污泥连续流反应器，探讨好氧反硝化细菌的投加对含硝氮废水处理效果的影响，结果表明强化系统对于硝酸盐的平均去除率高达98.35%。苏俊峰[85]将异养硝化细菌和好氧反硝化细菌投加至SBR反应器中，采用先硝化再反硝化的方式运行，反应器在稳定运行阶段，氨氮及TN的平均去除率分别为92.12%及74.15%，表现出较好的脱氮效能。张凯[86]将在水库中驯化筛选出的2株同步硝化反硝化细菌ZK-1和J×26及3株好氧反硝化细菌F4、DA15和HF6进行组合复配，制成微生态菌剂，直接投加到微污染水体中进行原位脱氮生物强化，与对照组相比，硝氮及TN的去除率均明显提高。魏清娟[87]将富集驯化的低温好氧反硝化菌群投加于强化系统中，反应器运行稳定后发现，相比于对照系统，强化系统出水TN去除率提高了约24.56%，强化系统的出水TN浓度能够达到《城镇污水处理厂污染物排放标准》[88]的一级A标准。张栋俊[89]将低温条件下驯化得到的好氧反硝化菌群投加至中试设备进行生物强化，强化后中试设备氨氮出水稳定在1 mg/L，TN出水稳定在10 mg/L，其中TN和氨氮分别比水厂提高了3.71%和23.18%。孙移鹿[90]将经菌丝球固定化的好氧反硝化菌T13投加至SBR反应器进行固定化生物强化，结果表明以O/A方式运行的SBR反应器的TN去除率高达77.0%；PCR-变形梯度凝胶电泳(PCR-denatured gradient gel electrophoresis，PCR-DGGE)结果表明，通过菌丝球的固定化作用，菌株T13可以长期稳定存留于系统内，形成优势种群，成功克服了游离态菌剂强化过程中易流失的问题；固定化载体不仅缓解了亚硝氮积累的问题，而且在弱酸和低温条件下对菌株T13均起到了一定的保护作用。王弘宇等[91]考察了固定于生物陶粒反应器上的好氧反硝化细菌对硝氮废水的处理效能，结果表明该反应器对硝态氮的去除率可以基本达到100%，PCR-DGGE结果显示，在整个运行阶段好氧反硝化菌在反应器中稳定存在，并且始终是优势菌群。
杨基先等[92]通过实验验证，磁场强度和磁作用方式均对好氧反硝化功能菌的生理特性产生影响，当磁场强度为150 mT时，硝氮去除率、脱氢酶活性分别较未加磁场时提高了7.2%和2.38倍。王强[93]将好氧反硝化菌T13接入SBR系统，考察生物-磁复合强化技术的低温同步脱氮效能及稳定性，结果表明低温运行期间，出水平均氨氮及TN去除率分别为95.76%和60%，出水氨氮浓度符合《城镇污水处理厂污染物排放标准》一级A标准，磁作用效果显著[88]。孙静文[94]将磁强化技术与好氧反硝化技术相结合，构建序批式生物膜反应器(sequencing biofilm batch reactor，SBBR)新型高效脱氮除碳工艺，SBBR平均出水氨氮和TN去除率分别为96.79%和76.19%，SBBR的总氮和COD去除效果及稳定性均优于对照系统，PCR-DGGE结果显示，好氧反硝化菌T13作为功能菌在运行过程中一直运作于磁强化载体内。
近年来，关于好氧反硝化菌在微污染水源水体中的报道逐渐增多。Su等[95]、Huang等[96-97]在微污染的水库水体中成功筛选出多株好氧反硝化菌，并探究其脱氮特性。Zhou等利用筛选出的本土好氧反硝化菌实现原位强化处理微污染水库水源水[14]。Wen等[41]发现天然有机物的分子量对好氧反硝化菌脱氮效能具有显著影响。Zhou等[77]发现春季的周村水库水体中氮素的损失主要是由好氧反硝化菌造成的。除了在水处理方面的应用，好氧反硝化菌也在烟气脱硝[98]、土壤修复等领域[99-100]广泛报道。
7 总结及展望好氧反硝化生物脱氮技术方面的研究如火如荼地进行了30余年，目前筛选的好氧反硝化菌种类繁多，在脱氮特性及功能基因表达方面的研究逐渐深入，但在中试及厂试规模的实际应用及好氧反硝化机理方面的研究仍显不足，存在以下几方面的局限性：
(1) 目前好氧反硝化生物脱氮技术在实际应用方面的研究中仍然停留在实验室小试阶段，中试及厂试规模的研究鲜有报道。主要原因在于不能有效控制好氧反硝化功能菌剂的流失，无法使其成为优势菌种。另一方面，后续研究即使是小试规模的实验，也应尽量使用实际废水进行实验。
(2) 关于好氧反硝化菌的动力学及化学计量学方面的研究鲜有报道。与传统反硝化菌相比，好氧反硝化菌的反硝化过程由于有氧气的参与，其反应动力学及化学计量学必然与传统反硝化菌的反应动力学及化学计量学存在不同。最近，Feng等[101]报道了一株好氧反硝化菌(P. stutzeri T13)的化学计量学及动力学方面的研究。好氧反硝化菌株的化学计量学系数以及动力学常数的获得，有助于对水处理工艺过程进行调控，使之适合好氧反硝化菌的生存，然而更多不同种属的好氧反硝化菌株需要进一步的试验。同时，关于异养硝化-好氧反硝化菌株的化学计量学及动力学方面的研究也需要进一步的探究。
(3) 未来的研究方向可以通过分子生物学的手段，通过基因编译等方法提高好氧反硝化菌胞内电子传递效率，提高好氧反硝化效能。另一方面，群体感应(quorum sensing，QS)现象是目前研究的热点之一，然而关于好氧反硝化菌的群体感应方面的研究还鲜有报道，未来可在此方向开展研究。
(4) 目前的研究内容主要针对的基本都是好氧反硝化菌的分离及生物脱氮特性研究，而针对好氧反硝化作用机理方面的研究需要进一步加强，尤其是在电子传递方面的研究有待进一步加强。
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表1(Table 1)
