2023, Vol. 44
2. 肇庆学院广东省环境健康与资源资源利用重点实验室, 肇庆 526061;
3. 中国科学院烟台海岸带研究所, 中国科学院海岸带环境过程与生态修复重点实验室, 烟台 264003;
4. 中国农业大学土地科学与技术学院, 北京 100193
2. Guangdong Provincial Key Laboratory of Environmental Health and Land Resource, Zhaoqing University, Zhaoqing 526061, China;
3. Key Laboratory of Coastal Environmental Processes and Ecological Remediation, Yantai Institute of Coastal Zone Research, Chinese Academy of Sciences, Yantai 264003, China;
4. College of Land Science and Technology, China Agricultural University, Beijing 100193, China
生物炭(biochar)是生物质在限氧条件下热解产生的一类难溶的、稳定的且高度芳香化的富碳固体物质, 因其具有孔隙结构发达、比表面积巨大、表面活性官能团含量丰富和阳离子交换量较大等特点, 可作为良好的阳离子重金属吸附材料, 并已应用于镉和铅等重金属污染土壤的稳定化修复[1, 2].但由于大多数生物炭表面带有负电荷, 对土壤中以砷酸盐和亚砷酸盐等阴离子形式存在的无机砷(As)吸附率较低[3, 4], 需要通过物理、化学或生物方法对生物炭材料进行改性或修饰, 以增强其对阴离子型污染物的吸附能力[5].
铁及其氧化物等铁基材料在环境中具有储量丰富和来源广泛, 具有表面电荷高、吸附能力强和易分离等优点, 对As具有良好吸附性能.通过化学法或物理法等方法提高生物炭中铁组分的含量(即进行铁修饰), 可以提高其对As的吸附固持能力, 如将纳米零价铁或磁性铁氧化物等颗粒负载在生物炭基底上, 能有效提高颗粒的分散性, 减缓颗粒表面钝化, 解决生物炭难以分离的缺点[6].铁修饰生物炭材料同时结合了生物炭与铁基材料二者的优良特性, 其在环境修复领域的应用研究正日益引起广泛重视.为增进对铁修饰生物炭理化特性及其与砷的反应机制的认识, 促进铁修饰生物炭在重金属污染土壤修复中的应用, 本文结合国内外相关最新研究进展, 综述了铁修饰生物炭材料的制备与表征方法, 阐明了铁修饰生物炭对砷的吸附过程与机制, 探讨了铁修饰生物炭对土壤中砷的稳定化效果与影响机制, 并对铁修饰生物炭的未来发展方向进行了展望.
1 生物炭和铁修饰生物炭材料的制备生物炭的生产是铁修饰生物炭材料制备的前提和基础.生物炭的生产涉及生物质原料类别、碳化方式和制备温度等多个方面.生物质材料来源广泛, 作物秸秆、园林垃圾、城市固废和粪便垃圾等均可以用来制备生物炭[7].不同材料的组成差异决定了生物炭的物质组成、孔隙结构、芳香性等基本性质[8~10].限氧热裂解法(主要为慢速热解法)是目前生物炭制备的主要方法[11].有研究发现, 随着热裂解温度的升高, 生物炭的含氧官能团减少, 比表面积和碳含量增大, 芳香性增强, 吸附能力有所提升[7, 12].
铁修饰生物炭的制备是通过在生物炭制备的过程中或制备完成后, 以铁盐溶液、铁粉/铁氧化物粉末或富铁微生物等形态加入铁基材料, 并在物理、化学和生物学的作用下与生物炭结合, 形成稳定的铁-炭复合材料.当前已报道的常见铁修饰生物炭合成方法包括沉淀法、热解法、球磨法和微生物改性法等(图 1)[13~16].
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图 1 铁修饰生物炭修复材料制备方法 Fig. 1 Preparation methods of iron-modified biochars |
沉淀法是目前应用较广泛的铁修饰生物炭制备方法之一, 包括共沉淀法和还原沉淀法.共沉淀法是较早被采用的铁修饰生物炭制备方法, 将生物炭分散在铁盐溶液中形成均质悬浊液, 用氢氧化钠(NaOH)等碱溶液调节pH至碱性(通常为9.0~11.0), 搅拌混合使生物炭表面和孔隙内形成铁氧化物沉淀, 经洗涤、干燥后获得铁修饰生物炭. Zhang等[17]发现原始杨木生物炭对As几乎没有吸附能力, 而采用表面共沉淀法制备氯化铁/赤铁矿复合修饰生物炭后, 铁修饰生物炭对As的吸附容量提高至3.15 mg·g-1. Kim等[18]采用NaOH调节硝酸铁-芒草生物炭的悬浊液pH值至9.0, 使炭表面形成无定型铁沉淀, 获得的铁修饰芒草生物炭对溶液中亚砷酸盐的去除率(>72.3%)显著高于原始生物炭(< 10.4%).朱司航等[19]用共沉淀法制备的纳米针铁矿修饰秸秆生物炭对溶液中As(Ⅲ)的最大吸附能力达65.20 mg·g-1, 相比原始生物炭提高了62.10倍.共沉淀法所需的合成设备简单, 成本相对较低, 易于实现工业化, 形成的铁氧化物稳定性好, 但需消耗大量溶剂和时间, 且沉淀形成过程较难控制[20, 21].还原沉淀法与化学沉淀法类似, 区别是在生物炭与铁盐溶液混合时需要加入硼氢化钠或硼氢化钾进行还原. Wang等[22]采用硼氢化钠液相还原法制备的纳米零价铁修饰生物炭对As(Ⅴ)的吸附容量最高可达124.5 mg·g-1.采用该方法制备的零价铁修饰生物炭对污染物通常有较好的去除效率, 但合成工艺较为复杂, 所需还原剂通常具有毒性, 形成的零价铁环境稳定性较差[14, 23].
1.2 热解法热解法主要包括浸渍热解法和直接热解法.浸渍热解法是将生物质在热解前先进行预处理, 用铁盐溶液[FeCl3、Fe(NO3)3等]或铁矿物(赤铁矿、针铁矿等)的悬浮液浸渍生物质原料, 使金属离子或铁矿物沉积在原料的表面与内部, 经低温烘干后再热解得到铁修饰生物炭[20, 24].通过铁盐浸渍联合高温热解的方式, 可使生物炭的比表面积增大, 生成大量的吸附位点, 以提高生物炭对目标污染物的吸附能力[25]. Zhou等[26]将栗壳生物炭、Fe2+/Fe3+混合液和明胶均匀搅拌, 于328 K下热解2 h得到磁性铁修饰生物炭.扫描电镜图片显示生物炭表面负载了大量不规则铁颗粒, 铁修饰后生物炭对As(Ⅴ)的吸附容量为45.8 mg·g-1, 高于未修饰生物炭(17.5 mg·g-1). Xu等[27]用杨木作为生物质原料, 经氯化铁(FeCl3)溶液浸泡后于300、600和900℃下热解, 得到的铁修饰生物炭比表面积均显著高于同温度下制备得到的未修饰生物炭, 分别为7.37、21.2和288 m2·g-1, 且对As(Ⅲ)的吸附容量为39.69、47.61和48.57 mg·g-1, 对As(Ⅴ)的吸附容量达87.33、121.02和121.61 mg·g-1.浸渍热解法操作简便, 但能耗较高, 且不能直接用于污泥等含水量高的生物质[15, 24].直接热解法是将含铁的生物炭原料(如污泥)直接热解, 或将生物质与含铁物质/溶液混合后直接热解制备铁修饰生物炭(水热碳化法)[21]. Cho等[28]以造纸厂污泥为原料, 在CO2氛围下热解制备的磁铁矿-碳酸钙(Fe3O4-CaCO3)修饰生物炭对As(Ⅴ)的吸附容量为34.1 mg·g-1. Bakshi等[6]将红橡木和柳枝稷生物质原料与天然磁铁矿悬浊液充分混合并烘干后, 在900℃ N2氛围下热解1 h得到的零价铁修饰生物炭, 对As(Ⅴ)的吸附量分别为6.48 mg·g-1和15.66 mg·g-1.
1.3 球磨法机械球磨技术是近年来兴起的一种新型改性生物炭制备技术, 是提升生物炭的物理化学性质和吸附性能的有效方法[29].机械球磨可将生物炭粒径降低至纳米级, 提高其内外表面积, 并暴露出更多的石墨烯结构和含氧官能团[30].球磨改性不仅是一种机械物理处理, 同时也是一种机械化学处理.球磨过程中生物炭的大分子化学键被破坏或拉伸, 从而增加了其表面官能团的种类和数量.通过与铁粉或铁氧化物一起研磨, 机械球磨还可以在生物炭表面引入纳米铁颗粒, 进一步提高生物炭的吸附性能[31].Shan等[32]采用行星式球磨仪制备的磁性Fe3O4修饰生物炭, 其比表面积和孔隙体积分别为365 m2·g-1和0.54 cm3·g-1, 较原生物炭有显著提升(30.9 m2·g-1和0.042 cm3·g-1), 对水中卡马西平和四环素的吸附能力可达62.7 mg·g-1和94.2 mg·g-1.球磨法也可与其他方法联用以制备超细铁修饰生物炭材料.如Li等[33]首先采用浸渍热解法将小麦秸秆制备成磁性生物炭, 进一步采用球磨法制备成纳米磁性生物炭后, 对汞的去除效率显著高于原始磁性生物炭.Zou等[34]采用浸渍热解法联合球磨法制备磁铁矿修饰竹炭(BM-Fe-HC), 发现球磨有效减小了BM-Fe-HC的粒径、提高其比表面积, 同时增加了Fe3O4在生物炭表面的附着量, 从而提高BM-Fe-HC对Cr(Ⅵ)的吸附速率与吸附容量.对于As(Ⅴ), Li等[35]通过干磨法制备的膨胀蛭石-核桃壳炭纳米复合材料比表面积和孔体积较原始炭增大了2~6倍, 复合材料表面的Mg2+和Fe3+等阳离子被活化, 含氧有机官能团增加, 对As(Ⅴ)的吸附能力也由球磨前的无吸附增加至20.1 mg·g-1. Zubrik等[36]将水基铁磁流体和热解褐煤生物炭在N2氛围下进行高能球磨, 得到的磁性铁修饰生物炭对As(Ⅴ)的最大吸附量为19.9 mg·g-1.最近, Yang等[37]将天然磁铁矿在空气氛围下进行球磨处理, 生成的纳米磁铁矿粒径和结晶度降低, 表面负载更多的Fe—OH和Fe—COOH官能团, 对水溶液中的As(Ⅲ)和As(Ⅴ)的吸附能力由球磨前的0.47 mg·g-1和1.14 mg·g-1分别提高至6.95 mg·g-1和3.16 mg·g-1.总的来说, 当前有关球磨法制备铁修饰生物炭及其在砷的吸附去除方面的研究仍非常有限, 未来仍需进一步开展深入研究.
1.4 (微)生物改性法等其他铁修饰方法生物法是将某些具有铁富集功能的微生物与生物炭结合以改善生物炭表面性质的方法.Luo等[38]首次通过培养黄孢原毛平革菌(Phanerochaete chrysosporium)将铁离子富集于菌体中, 作为富铁生物质原料, 在700℃氩气氛围中热解得到多孔磁性生物炭材料.该磁性生物炭具有较大的比表面积(1 986 m2·g-1), 对双氯芬酸的去除能力较好(361.25 mg·g-1), 同时易于固液分离.尽管生物法是一种节能环保的生物炭修饰方法, 但仍存在一些不足之处, 如微生物产量有限和改性速率较慢等.此外, 也有研究者采用微波法[39]和电磁法[40]等方法制备铁修饰生物炭, 但均处于实验室探索阶段, 尚无具体应用案例.表 1总结了目前铁修饰生物炭主要制备方法的优缺点.在选择铁修饰生物炭的制备方法时, 应充分考虑每种方法的适用条件与限制因素, 以及环境中污染物的总量及形态特征、土壤理化性质、时间、成本和易操作性等多种因素进行综合选择.
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表 1 常见铁修饰生物炭材料制备方法的优缺点 Table 1 Advantages and disadvantages of the common preparation method of iron-modified biochar materials |
2 铁修饰生物炭对砷的吸附机制
铁修饰生物炭结合了生物炭与铁基材料两者独特的理化性质, 对污染物的吸附去除性能常常优于单一材料.已有研究通过采用一种或多种分析技术, 对铁修饰生物炭材料的理化性质和结构特征进行表征, 并研究材料对砷的吸附机制(表 2).一般来说, 铁修饰生物炭对环境中砷的作用机制主要包括3个方面：生物炭基底和铁修饰材料分别对砷的作用机制, 以及两者对砷的协同作用机制.
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表 2 铁修饰生物炭材料的表征技术及其对砷吸附机制的研究方法 Table 2 Characterization techniques of iron-modified biochar material and research methods for its arsenic adsorption mechanism |
生物炭对As的吸附机制包括表面络合、离子交换、静电作用、氧化还原、(共)沉淀和孔隙扩散等[61].如生物炭表面的羟基、羧基等含氧官能团可通过表面络合、静电作用和离子交换反应与As结合; 而羧基和酚羟基也可作为电子供体, 将As(Ⅴ)还原为As(Ⅲ).与此相反, 生物炭中的一些原生矿质组分如纤铁矿[γ-FeO(OH)]可促使As(Ⅲ)转变为As(Ⅴ)并吸附, 从而降低其环境风险[62, 63].此外, 生物炭中的盐基离子如Ca2+和Mg2+可与As(Ⅲ)/As(Ⅴ)形成Ca(AsO2)2、CaAsO2OH·H2O和Ca5(AsO4)3OH, 以及Mg(HAsO4)2和(Mg3(AsO4)2等沉淀并沉积在生物炭孔隙结构中[64, 65].上述作用机制中, Li等[66]认为表面络合和静电作用是主要的吸附机制, 而生物炭对As的氧化还原和沉淀作用微弱.Lee等[67]研究表明, 在由不同温度下制备的生物炭对As的吸附过程中, 10%~64%的As(Ⅴ)可被还原为As(Ⅲ)和As(0), 再通过物理吸附和疏水作用被固持, 而剩下的90%~36%的As(Ⅴ)主要通过与阳离子形成静电桥接结构被吸附.鉴于As(Ⅴ)的还原产物As(Ⅲ)具有更高的环境迁移性和生物毒性, 因此, 由生物炭引发的As(Ⅴ)还原应当引起重视.
生物炭上的铁基修饰材料(如铁氧化物、纳米零价铁和硫化亚铁等)对砷的吸附机制则包括静电作用、表面络合、配位离子交换、氧化还原和(共)沉淀等[68, 69].Cho等[24]采用FeCl3溶液浸泡咖啡渣, 分别置于CO2和N2氛围下热解, 得到的生物炭表面的主要铁成分分别为磁性Fe3O4和碳化铁(Fe3C).尽管前者的比表面积比后者大70倍, 但对As(Ⅴ)的吸附量(13.1 mg·g-1)却低于后者(8.9 mg·g-1), 这主要是由于Fe3C的等电点(pHZPC=10)较Fe3O4的(pHZPC=7.5)更高, 表面负载正电荷更多, 从而易于通过静电引力的作用带与砷酸根离子亲合.Feng等[59]采用同步辐射X射线吸收谱学技术发现, 磁性铁修饰生物炭表面负载的铁组分主要包括磁赤铁矿(γ-Fe2O3)、Fe0、铁硫化物(FeS1-x)和少量残留的FeCl3, 其中γ-Fe2O3等铁氧化物能够通过络合作用与As(Ⅴ)或As(Ⅲ)形成稳定的内层单齿或双齿配合物, 进而降低As的活性和生物有效性; 而Fe0溶解释放出的Fe2+, 可与As发生共沉淀而被固持在材料表面.在这项研究中还观测到有81.5%的As(Ⅲ)被氧化为As(Ⅴ), 提示铁基材料不仅可以提高对As的吸附固持能力, 还可以通过氧化还原作用促进高毒高迁移性的As(Ⅲ)向低毒低迁移性的As(Ⅴ)转化, 减小其环境风险.
铁修饰生物炭中, 生物炭基底及其铁修饰材料除了可分别对砷进行吸附作用, 其铁炭复合体也可协同发挥固砷作用, 主要包括3个方面：①生物炭作为基底, 能有效增加铁组分的分散性及其与As的接触面积; ②生物炭通过自身吸附过程, 加速As从环境介质中向复合材料表面的传质过程, 提高铁修饰材料与As的反应速率[70]; ③Fe0等铁组分与炭基底形成的铁碳原电池, 可提高材料As的去除能力, 并可解决Fe0在反应过程中生产惰性层或金属氢氧化物导致去除率降低的缺陷[71, 72].Yu等[73]以两种含铁污泥为前体材料, 分别在350℃和700℃条件下制备了具有不同Fe含量和种类的铁修饰生物炭(SSA和SSB).两者的Fe含量分别为12.7%~17.7%和3.6%~8.6%, 而SSA对As(Ⅴ)的吸附容量可达60.2~90.2 mg·g-1, SSB的仅为0.03~1.06 mg·g-1.这主要是由于SSB生物炭基底带负电荷, As(Ⅴ)因静电互斥作用难以同富铁点位接触发生反应.而SSA生物炭基底带正电荷, 有利于As(Ⅴ)发生吸附, 并且不同温度条件下制备的SSA材料均可释放出Fe2+, 可直接形成Fe(Ⅱ)-As(Ⅴ)共沉淀, 或经氧化沉淀形成Fe(Ⅲ)-水合氧化物后再对As(Ⅴ)进行吸附.笔者前期研究结果显示, 通过化学沉淀法制备得到铁碳质量比为1∶2的水铁矿修饰梨木生物炭对As(Ⅲ)的最大吸附量为47.6 mg·g-1, 而在相同铁炭比下, 相应质量的纯水铁矿与原始梨木生物炭对As(Ⅲ)吸附量分别为28.2 mg·g-1和0.25 mg·g-1, 说明铁炭复合达到了“1+1>2”的吸附效果.图 2总结了铁修饰生物炭对砷的主要吸附机制.
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图 2 铁修饰生物炭对砷的吸附机制 Fig. 2 Likely mechanisms of arsenic adsorption onto iron-modified biochar |
稳定化修复技术是通过向土壤中添加稳定化药剂, 经吸附、沉淀、络合、离子交换和氧化还原等一系列反应降低重金属的生物有效性和迁移性, 从而减少对动植物的毒性.该修复方法因具有能耗低、修复快速、操作简单等优越性, 能更好地满足当前我国重金属污染土壤低碳高效修复的迫切需求.作为一种稳定化修复功能材料, 生物炭和铁修饰生物炭已开始用于重金属污染土壤的稳定化修复[74~77].
基于生物炭对砷的吸附特性, 原始生物炭对砷污染土壤的稳定化修复作用通常有限.生物炭富含醌基等氧化还原活性结构, 可通过非生物或生物过程促发土壤原生铁氧化物与含砷矿物的解离和次生矿物的生成, 从而影响As的释放、再分配特征及赋存形态[78, 79].El-Naggar等[80]研究发现, 施入稻秆生物炭可促使矿区重金属复合污染土壤中的臭葱石(FeAsO4·2H2O)发生溶解并形成不稳定的氧化铁结合态As(如水铁矿结合态和针铁矿结合态), 并导致土壤中的As(Ⅴ)还原为As(Ⅲ).铁修饰生物炭往往较生物炭具有更好的砷稳定化效果.Wu等[74]向砷污染水稻土中分别施加了水稻秸秆生物炭、FeOS修饰生物炭、FeCl3修饰生物炭以及Fe0修饰生物炭, 在稳定化培养90 d后, 较无添加炭材料的对照土壤, 施加生物炭处理土壤中碳酸氢钠提取态As含量略有升高, 而FeOS修饰生物炭、FeCl3修饰生物炭和Fe0修饰生物炭处理土壤中易释放态As含量分别下降了13.95%~30.35%、10.97%~28.39%和17.98%~35.18%.在对重污染场地土壤(Cr、Cu和As的总量分别为540、1 662和1 364 μg·g-1)的稳定化修复研究中, 生物炭和零价铁联合处理将土壤中水提取态Cr、Cu和As的含量分别降低了45%、45%和43%, 并使土壤细菌群落结构多样性得到有效恢复[81].
铁修饰生物炭施用于真实的砷污染土壤时, 其铁/炭组分可能会受到Eh和pH等土壤环境条件的影响而发生形态转变, 进而影响对土壤中砷的释放和稳定化修复效果.Yin等[82]报道了秸秆生物炭(pH 10.7)的添加提高了土壤pH, 进而提高了土壤孔隙水中As的溶解性; 而氯化亚铁(FeCl2)修饰生物炭(pH 4.87)的添加在未显著改变土壤pH的情况下, 降低了根际土壤中As的溶解度, 抑制了As的释放.Yang等[83]采用生物地球化学微宇宙试验装置, 研究了生物炭和铁修饰生物炭在不同氧化还原条件下对As的固定效应.对于未施炭的砷污染水稻土对照处理, 在Eh从200 mV降低至-300 mV的过程中, 土壤溶液中ρ(As)从检出限以下迅速升至1 535.8 μg·L-1; 而在随后的曝氧期, 即Eh从0 mV升至300 mV过程中, ρ(As)从686.2 μg·L-1逐渐降至检出限以下.与此同时, 施加生物炭处理组土壤的As释放量较对照土壤降低了16.0%~41.3%; 而施加铁修饰生物炭的土壤较对照土壤降低了32.6%~81.1%的As释放量.在淹水条件下, 添加生物炭, 尤其铁修饰生物炭, 可以通过形成稳定的As(Ⅲ)-水铁矿、As(Ⅴ)-水铁矿和As(Ⅲ)-三价铁氧化物-腐殖质复合物等铁砷有机络合物固定土壤中的As[83].Fan等[84]研究发现, 采用纳米零价铁修饰生物炭(nZVI/BC)改良矿区As污染土壤时, 受土壤中氧化作用以及Ca2+和Al3+等阳离子影响, 土壤中硫酸铵提取态As的含量降低了93%以上, 模拟胃肠提取态As含量降低了85%以上.最近, Zhang等[85]评估了干湿交替和冻融交替条件下零价铁修饰生物炭(ZVI/BC)对砷污染土壤的稳定化修复效果, 发现干湿交替可导致ZVI的氧化溶解和As(Ⅲ)的形成, 使土壤中ω(生物有效态As)从1.25~9.50 mg·kg-1增加到1.83~21.75 mg·kg-1; 而冻融交替可促进土壤中无定形铁向结晶形铁转化, 使As(Ⅲ)氧化为As(Ⅴ), 进而导致土壤中ω(生物有效态As)从9.50~1.25 mg·kg-1降低至5.42~0.45 mg·kg-1.此外, 铁修饰生物炭还可通过增加土壤中As/Fe还原菌的相对丰度调控As(Ⅴ)/Fe(Ⅲ)的还原[75].综上, 铁修饰生物炭修复后土壤中As的稳定性仍会随着不同老化过程而发生变化, 这需要综合考虑复杂环境条件中非生物与生物等多因素的影响.当前, 对于铁修饰生物炭在砷污染土壤的稳定化修复中的应用研究仍非常有限, 关于铁修饰生物炭调控As在土壤和生物炭界面的键合形态和固持与释放的边界条件等关键科学问题尚未有明确的结论.仅有的少量研究是在实验室内进行的盆栽试验, 鲜见在砷污染的农田或场地土壤中开展实际应用与示范推广.
4 展望(1) 将农林废弃物资源化利用与生物炭制备相结合, 继续研发高效稳定、成本低廉、低碳环保的铁修饰生物炭产品, 进一步提高铁修饰生物炭对污染物的吸附去除性能.结合智慧农业与人工智能技术, 创新研发铁修饰生物炭材料在污染农田或场地土壤(地下水)修复过程中的精准靶向施用或注射装备, 推进应用示范与产业化发展.
(2) 采用多种先进的同步辐射谱学和成像技术, 结合同位素示踪和多组学技术(如代谢组、基因组、宏基因组), 阐明铁修饰过程对生物炭基组分和结构对砷吸附性能的影响, 在分子层面精准剖析铁-生物炭复合体系对污染土壤中砷的键合特征与稳定化机制, 探索铁修饰生物炭在固砷同时的固碳潜力及其协同作用机制, 服务于重金属污染土壤的减污降碳协同增效修复.
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