Adsorption behavior of copper ions on graphene oxide–chitosan aerogel

第42卷第2期青岛科技大学学报(自然科学版)2021年4月Journal of Qingdao University of Science and Tcchnology(Natural Science Edition)Vol..2N o. Apr.2021文章编号:1672-6987(2021)02-0058-08；DOI：10.16351/j.1672-6987.2021.02.008乙二胺四乙酸(EDTA)改性磁性壳聚糖对Cd2+的吸附性能于硕，吴占超*,匡少平*(青岛科技大学化学与分子工程学院，山东青岛266042)摘要：成功制备了乙二胺四乙酸(EDTA)改性的磁性壳聚糖，并通过红外光谱,X射线衍射，热重分析和扫描电镜对其结构和形貌进行了表征。

对吸附剂的吸附性能研究表明：在p H=5,T=298K,p0=200m g-L1t=30min的吸附条件下，吸附剂对Cd2的饱和吸附量为176.32mg・g1.吸附剂吸附行为符合二级动力学模型和Langmuir等温模型。

吸附剂再生5次仍有较好的吸附性能。

关键词：磁性壳聚糖；乙二胺四乙酸(EDTA)；Cd2；吸附中图分类号：O646.8文献标志码：A引用格式：于硕,吴占超,匡少平.乙二胺四乙酸(EDTA)改性磁性壳聚糖对Cd2的吸附性能青岛科技大学学报(自然科学版)，2021,42(2)：5865.YU Shuo,WU Zhanchao,KUANG Shaoping.Adsorption properties of ethylenediamine tet­raacetic acid(EDTA)modified magnetic chitosan for Cd2[J].Journal of Qingdao Universi­ty of Science and Technology(Natural Science Edition),2021,42(2)：5865.Adsorption Properties of Ethylenediamine Tetraacetic Acid(EDTA)Modified Magnetic Chitosan for Cd2+YU Shuo,WU Zhanchao,KUANG Shaoping(College of Chemistry and Molecular Engineering,Qingdao University of Science and Technology,Qingdao266042,China)Abstract:EDTA-modified magnetic chitosan was successfully prepared,and its structure andmorphologywerecharacterizedbyinfraredspectroscopy,X-raydi f raction,thermogravimet-ricanalysis,andscanningelectron microscopy.Thestudyontheadsorptionperformanceofthe adsorbent,showed that,the saturated adsorption capacity of the adsorbent,for Cd2was176.32mg・g1under the adsorption conditions of pH=5,T=298K,^0=200mg・L1and t=30min.The adsorption behavior of the adsorbent,complies with the second-order ki-neticmodelandtheLangmuirisothermalmodel.Theadsorbentsti l showedgoodadsorptionpropertyafterthefifthregenerationcycle.Key words:magnetic chitosan；ethylenediamine tet.raacet.ic acid(EDTA)；C d2；adsorption现代工业领域产生的有毒金属离子对水体的污多的有毒金属中，Cd?被认为是最剧毒的一种。

※个人简介刘转年，男，陕西富平人，博士（后），教授，博士生导师。

本科毕业于南京理工大学环境工程专业，2004年博士毕业于西安建筑科技大学环境工程专业。

2006-2009年在西安建筑科技大学材料科学与工程博士后流动站从事博士后研究，2009年破格晋升教授。

2014年7月至12月美国Northern Arizona University 访问学者。

西安科技大学环境科学与工程一级学科硕士点召集人，环境工程二级学科硕士点带头人。

高等学校（矿业）“十二五”规划教材环境学科编审委员会委员。

国家自然科学基金项目同行评议专家、中国博士后科学基金项目评审专家、陕西省博士后基金项目评审专家，西安节能协会特聘专家。

※研究方向1、矿物环境功能材料及应用研究；2、矿井水处理及资源化利用新技术；3、矿业固体废物资源化利用。

※主要成果在国内外核心期刊发表论文60余篇，其中SCI和EI收录20余篇，出版专著和教材4部。

主持国家自然科学基金面上项目、陕西省工业攻关项目、中国博士后科学基金项目、西安市工业攻关项目、省教育厅产业化培育项目、榆林市产学研合作项目及西安市发改委科研课题等纵、横向项目近20项。

获省部级科学技术二等奖和三等奖各1项，个人排名第一。

2015年获中国化工学会和化工学报高被引论文奖。

授权国家发明专利12项。

代表性成果：[1]刘转年著.粉煤灰成型吸附剂的制备及应用[M].北京:化学工业出版社,2009.[2]刘转年等编著.环境污染治理材料[M].北京:化学工业出版社, 2013.[3]刘转年,范荣桂.环保设备基础[M].徐州:中国矿业大学出版社, 2013.[4]Liu Zhuannian, Zhang Yuanyuan, An Yangkang, Jing Xiuyan, Liu Yuan. Influence of coal fly ash particle size on structure and adsorption properties of forming adsorbents for Cr6+ [J].Journal Wuhan University of Technology, Materials Science Edition, 2016, 31(1):58-63. SCI检索.[5]Zhuannian Liu,Yuan Liu.Structure and properties of forming adsorbents prepared from different particle sizes of coal fly ash[J]. Chinese Journal of Chemical Engineering,2015,23(1):290-295. SCI检索.[6]Zhuannian Liu, Yongmei Liu, Long Chen, Huan Zhang. Performance study of heavy metal ion adsorption onto microwave- activated banana peel [J].Desalination and Water Treatment, 2014,52 (37-39):7117-7124. SCI检索.[7]Liu Zhuannian, Song Yejing, Han Xiaogang. Synthesis and characteristics of a novel heavy metal ions chelator[J].Journal Wuhan University of Technology, Materials Science Edition. 2012,27(4): 730-734. SCI检索.[8]Liu Zhuannian, Zhou Anning, Wang Guirong.Adsorption behavior of methyl orange onto modified ultrafine coal powder [J]. Chinese Journal of Chemical Engineering, 2009,17(6):942-948. SCI检索.[9]Liu Zhuannian, Wang Guirong. Removal of Cr (VI) from aqueous solution using ultrafine coal fly ash[J].Journal Wuhan University of Technology, Materials Science Edition, 2010, 25(2): 323-327. SCI检索.[10]Liu Zhuannian. Experimental research on oxidation of phenol by activated persulfate [J]. Journal of Coal Science & Engineering, 2013,19(4):560-565.[11]刘转年,张焕,王贵荣,程爱华,王艺.煤基重金属螯合吸附剂的制备及性能研究[J].煤炭学报,2015,40(1):172-178. EI检索.[12]刘转年,刘源.超细粉煤灰基成型吸附剂的制备及性能研究[J].煤炭学报,2009, 34(9):1263-1267. EI检索.[13]刘转年,杨志远.超细粉煤灰吸附Cr6+机理和动力学[J].中国矿业大学学报, 2008,37(4):478-482. EI检索.[14]刘转年.一种以乙二胺为原料的煤基复合螯合材料的制备方法.中国发明专利,ZL201410172439.2, 2016-01-05.[15]刘转年,王贵荣,王艺.一种聚合硫酸铁的制备方法.中国发明专利, ZL201310215351.X , 2015-04-22.[16]刘转年.一种煤基复合螯合剂的制备方法,中国发明专利, ZL201210593983.5, 2014-04-30.[17]刘转年.一种煤基螯合吸附剂的制备方法.中国发明专利, ZL201210594048.0,2014-07-23.[18]刘转年,刘永梅.一种纳米Fe0/多乙烯多胺复合螯合剂的制备方法.中国发明专利, ZL20131021376 9.7, 2015-10-14.[19]刘转年,陈龙.一种煤/聚乙烯亚胺交联复合螯合吸附剂的制备方法.中国发明专利, ZL201410039 800.4, 2015-12-02.[20]刘转年.一种重金属螯合吸附剂的制备方法.中国发明专利, ZL201310073432.0, 2015-12-02.[21]刘转年,刘永梅.一种纳米Fe0/聚乙烯亚胺复合螯合剂的制备方法.中国发明专利, ZL20131021376 7.8, 2015-12-09.在研科研项目：1、国家自然科学基金面上项目：煤/聚乙烯亚胺交联复合螯合吸附剂的制备及其对重金属离子的协同作用机制研究，项目编号：51278418。

第52卷第9期 辽 宁 化 工 Vol.52，No. 9 2023年9月 Liaoning Chemical Industry September，2023收稿日期： 2022-09-15Cu 2+在醋酸盐离子液体中的电化学性能及电沉积于开鑫1，刘艳辉1,*，宋继梅2,*，罗万胜1（1. 沈阳理工大学 材料科学与工程学院，辽宁 沈阳 110159；2. 潍坊科技学院 化工与环境学院，山东 潍坊 262700）摘 要：采用循环伏安法（CV）研究了二价铜离子在1-丁基-3-甲基咪唑醋酸盐[C 4C 1Im][OAc]离子液体中的氧化还原过程及电化学行为。

实验结果表明：[C 4C 1Im][OAc]离子液体的电化学窗口为3.3 V；铜离子在[C 4C 1Im][OAc]中的氧化还原为非可逆过程，铜离子还原过程受扩散传质控制，由Cu 2+→Cu +、Cu +→Cu 0的扩散系数分别为0.000 939 4，0.001 752 cm 2/s, SEM 及XRD 分析表明铜离子在醋酸盐离子液体中可以被沉积出来。

关 键 词：电化学窗口；离子液体；醋酸；传质机制；电沉积中图分类号：TQ153.14 文献标识码： A 文章编号： 1004-0935（2023）09-1306-04铜具有很好的导电、导热、延展性及机械加工性能，常运用于印刷电路板等电子工业领域[1-3]。

铜的电解冶炼具有悠久的发展历史，目前常用的电沉积铜体系主要包括水相硫酸盐体系、水相焦磷酸盐体系、氰化物体系以及无氰镀铜体系等[4]，但这些体系存在着许多缺点，例如工艺过程复杂、能源使用效率低、环保压力大、沉积层的质量较难控制等[5]。

离子液体又叫做室温熔盐、有机离子液体（ILs），作为一种绿色、安全的溶剂，由于其具极低的蒸汽压、高热稳定性、宽电化学窗口和高导 率[6]，越来越受到电化学家的欢迎。

自2013年，Liu 等[7]在1-乙基-3-甲基咪唑乙基硫酸盐离子液体中电沉积出具有纳米级别的微观结构的铜以来，铜的非水体系电化学性质已有大量的研究，目前已报道的电沉积铜及其合金的离子液体电解液体系主要有以下几种：氯化胆碱[8]、咪唑类四氟硼酸盐[9]、咪唑类三氟甲磺酸盐[10-11]、咪唑六氟磷酸盐等[5]。

第42卷第4期2023年4月硅㊀酸㊀盐㊀通㊀报BULLETIN OF THE CHINESE CERAMIC SOCIETYVol.42㊀No.4April,2023水滑石复合水泥基材料氯离子吸附能力的研究进展周丽娜1,2,蔡㊀颖1,马财龙1,罗㊀玲1,2(1.新疆大学建筑工程学院,乌鲁木齐㊀830017;2.新疆土木工程技术研究中心,乌鲁木齐㊀830017)摘要:以氯离子为诱导的钢筋锈蚀是造成混凝土耐久性问题的主要原因,其本质是氯离子通过材料基体的多孔结构在水泥基材料中扩散,并逐步迁移到钢筋表面,发生不利的物理化学反应㊂水滑石即层状双金属氢氧化物(LDHs)是一种新型延缓钢筋锈蚀的外掺材料,具有独特的层状结构和离子交换性质,可在特定的介质溶液中将客体阴离子与层间阳离子进行交换,达到吸附氯离子㊁延长混凝土结构服役寿命的目的㊂本文介绍了水滑石的结构性质㊁制备方法及氯离子吸附机理,总结了不同类型水滑石的氯离子吸附能力及相关研究成果㊂研究结果表明:水滑石复合水泥基材料的氯离子吸附性能受LDHs材料制备工艺㊁水泥基材料中孔隙液pH值及氯离子浓度影响,高温焙烧处理的水滑石对氯离子吸附效果更好;当LDHs掺量控制在1%~3%(质量分数)时,有利于改善水泥基材料的抗氯离子渗透性能㊂关键词:氯离子侵蚀;水滑石;水泥基材料;氯离子吸附;钢筋锈蚀;离子交换中图分类号:TU528㊀㊀文献标志码:A㊀㊀文章编号:1001-1625(2023)04-1137-11 Research Progress on Adsorption Capacity of Hydrotalcite forChloride Ions in Cement-Based MaterialsZHOU Lina1,2,CAI Ying1,MA Cailong1,LUO Ling1,2(1.School of Civil Engineering and Architecture,Xinjiang University,Urumqi830017,China;2.Xinjiang Civil Engineering Technology Research Center,Urumqi830017,China)Abstract:Chloride ions induced corrosion of reinforcing steel is a major cause of concrete durability problems.The essenceis that chloride ions diffuse through the porous structure of the material matrix in the cementitious material and gradually migrate to the surface of the reinforcing steel,where adverse physicochemical reactions occur.Hydrotalcite,known as layered double hydroxides(LDHs),is a new type of admixture to delay the corrosion of reinforcement.It has unique layered structure and ions exchange properties which ensures ion exchange between chloride ions and interlayer cations of LDHs and reaches the purpose of adsorption of chloride ions,thus extending the service life of concrete structures.This paper introduces the structural properties,modification method of hydrotalcite as well as its adsorption mechanisms on chloride ions,and summarizes the research results of recent years on the adsorption capacity of chloride ions through different modified hydrotalcite systems to improve the corrosion of steel reinforcement.The results indicate that the chloride ions adsorption capacity of hydrotalcite-based composite materials is mainly affected by LDHs preparation procedure,pH value of the pore solution and the chloride ions concentration.The chloride ions adsorption of calcined LDHs is more effective.When the addition of LDHs is1%~3%(mass fraction),the resistance to chloride ions permeability of cement-based materials will be obviously improved.Key words:chloride ions erosion;hydrotalcite;cement-based material;chloride ion adsorption;steel corrosion;ion exchange㊀收稿日期:2022-11-24;修订日期:2023-01-18基金项目:新疆维吾尔自治区高层次人才引进项目(TCBR202107);天山青年计划项目(2020Q071);新疆维吾尔自治区青年基金(2020D01C057)作者简介:周丽娜(1985 ),女,博士,讲师㊂主要从事混凝土材料耐久性损伤监测与评估的研究㊂E-mail:linazhou@通信作者:马财龙,博士,副教授㊂E-mail:macailong@1138㊀水泥混凝土硅酸盐通报㊀㊀㊀㊀㊀㊀第42卷0㊀引㊀言混凝土是建筑工程中广泛应用的材料之一,其耐久性能备受国内外学者关注㊂因混凝土自身的孔隙结构,外界环境中的氯离子通过孔隙渗入到混凝土内部,当氯离子浓度在钢筋表面达到一定阈值,钝化膜将会发生破坏,诱发钢筋锈蚀[1],严重影响混凝土结构服役寿命[2-5]㊂为了延缓腐蚀的发生,国内外众多学者和工程师致力于钢筋防腐技术的研究,目前相对成熟的保护技术包括阴极保护㊁防腐涂层㊁镀锌钢筋㊁缓蚀剂等[6]㊂上述方法对于延缓钢筋锈蚀均能发挥一定作用,但存在成本高㊁有毒性㊁难降解等问题㊂因此,需要开发一种直接㊁高效且可持续发展的技术㊂近年来,水滑石(layered double hydroxides,LDHs)因其独特的层状结构和离子交换性质,已成为一种新型的外掺材料,应用于混凝土结构中延缓钢筋腐蚀[7-8]㊂2003年Tatematsu 等[9]最早提出类似钙铝类水滑石材料可作为盐类吸附剂用于水泥基材料中,Raki 等[10]进一步证明类水滑石相材料掺入混凝土可控制有机混合物的释放速率㊂基于此,国内外学者从改性的角度出发,探究LDHs 对水泥基材料氯离子吸附性能的影响㊂在钢筋防腐方面,Hong 等[11]和Chen 等[12]通过试验研究表明LDHs 在含氯化物的模拟混凝土孔隙液中表现出良好的耐腐蚀性能,前者设计的MgAl-LDHs 膜用于钢筋基体上,能有效抑制钢筋锈蚀,后者制备的Ca-Al-NO 3LDHs 改性胶凝涂层具有固结氯离子和释放缓释阴离子的双重能力㊂上述研究表明,LDHs 固化氯离子效果显著㊂本文从LDHs 的结构性质和形成机制出发,引入LDHs 的氯离子吸附机理,考虑了LDHs 类型㊁制备方法及水泥基材料孔隙溶液中的pH 值㊁氯离子浓度等主要影响因素,对国内外学者针对不同类型LDHs 复合水泥基材料吸附氯离子能力的研究成果进行归纳梳理,以期为LDHs 在延缓混凝土结构钢筋锈蚀的运用提供理论依据,为延长钢筋混凝土建筑物和构筑物服役寿命提供新思路㊂1㊀水滑石复合水泥基材料1.1㊀水滑石的结构和性质图1㊀LDHs 结构示意图[15]Fig.1㊀Structure schematic diagram of LDHs [15]LDHs 是具有典型层状结构的纳米材料,其化学通式为[M 2+1-x M 3+x (OH)2][A n -]x /n ㊃y H 2O,其中M 2+和M 3+分别表示层板结构中的二价和三价金属阳离子,常见的二价金属阳离子M 2+包括Mg 2+㊁Mn 2+㊁Fe 2+㊁Co 2+等,常见的三价金属阳离子M 3+包括Cr 3+㊁Fe 3+㊁Mn 3+㊁Ga 3+等,A n -为层间的n 价无机(有机)阴离子,如NO -3㊁CO 2-3㊁Cl -㊁OH -㊁SO 2-4等,x 是M 3+/(M 2++M 3+)的摩尔比,一般在0.20~0.33;y 是每个水滑石分子中结晶水的个数[13-14]㊂典型的LDHs 结构[15]如图1所示㊂LDHs 具有离子交换特性,因此也被称为阴离子黏土[16]㊂在钢筋防腐方面,层间阴离子可以与外部环境中存在的侵蚀阴离子置换㊂若预先制备插层阴离子具有缓蚀效果的LDHs,还可在捕捉水泥基材料中游离氯离子的同时,释放抑制腐蚀的阴离子,发挥LDHs 的 双重效应 [17-18]㊂在进行层间交换时,一般需遵循层间阴离子的亲和力顺序:CO 2-3>SO 2-4>OH ->F ->Cl ->Br ->NO -3[19]㊂基于热稳定性和记忆效应,焙烧改性处理后的LDHs 置于特定介质溶液中,可以完成结构重建[20]㊂1.2㊀水滑石的制备水泥熟料中铝酸三钙(C 3A)和铁铝酸四钙(C 4AF)可促进水化产物单硫型水化硫铝酸钙(AFm)的形成,进而完成对氯离子的化学吸附[21-22]㊂AFm 相类属水滑石家族,证实了LDHs 复合水泥基材料使用的可行性[23]㊂水泥中活性MgO 和Al 2O 3的相对含量是形成LDHs 的关键[24],除了天然存在形式,水泥基材料中的LDHs 还可通过外掺的方式获得,根据水泥基材料的性能要求选择合适的掺量㊂外掺LDHs 常用制备工艺包括四种方法:共沉淀法㊁离子交换法㊁水热法和焙烧还原法㊂第4期周丽娜等:水滑石复合水泥基材料氯离子吸附能力的研究进展1139㊀1)共沉淀法共沉淀法为最常用的制备方法,即在一定温度下,将两种金属盐的混合溶液加入碱性溶液中,发生共沉淀以制备LDHs [25-26]㊂朱清等[27]以沉淀剂和温度作为制备MgAl-LDHs 的变量,结果显示,采用NaOH 为沉淀剂㊁晶化温度为70ħ时制备出的MgAl-LDHs 结晶度较好,晶相单一,晶体结构一致㊂Cao 等[28]采用共沉淀法成功制备了CaAl-LDHs,结果表明不同含量的CaAl-LDHs 对硅酸三钙(C 3S)和铝酸四钙(C 4A)的水化速度影响不同,进而使得浆体早强性能不同㊂这一发现为LDHs 在水泥基材料中的实际应用提供理论支持㊂2)离子交换法离子交换法主要基于LDHs 层间离子交换 的特性,根据离子的亲和力大小,进行主客体交换,以得到特定的阴离子插层的LDHs [29]㊂值得注意的是,用于制备盐溶液中的阴离子(即目标阴离子)亲和力应大于原有层间阴离子㊂离子交换法的优点在于反应时间短,是合成非CO 2-3插层LDHs 的常用方法㊂图2为离子交换法示意图㊂图2㊀离子交换法示意图[15]Fig.2㊀Schematic diagram of ion exchange method [15]3)水热法水热法是一种湿化学方法,在密闭压力容器内完成,利用高压㊁高温的水溶液使不溶或难溶的物质通过溶解或反应生成该物质的溶解产物,使其呈过饱和态,进而使得结晶生长[30]㊂水热法的优点是,制备过程在封闭环境中进行,反应产物受到大气环境中CO 2影响较小,产物结晶度好[31]㊂Chen 等[32]和杨成梅等[33]通过水热法,分别合成以硝酸根(NO -3)为插层的LDHs 和Ca-Al-Cl LDHs㊂前者研究了不同LDHs 的氯离子吸附规律,后者研究发现,复配比例下改性的LDHs 结构紧密,提高了水泥注浆早强性能㊂4)焙烧还原法焙烧还原法是根据LDHs 的记忆效应建立的制备方法㊂其基本原理是,通过对LDHs 进行高温煅烧,得到焙烧改性水滑石(calcined layered double hydroxides,CLDHs),将CLDHs 置于特定的介质溶液中完成结构重建[34]㊂焙烧还原法的优点是,在煅烧过程中脱除CO 2-3,提高客体阴离子的插层率㊂图3为CLDHs 层状结构的重建过程㊂图3㊀CLDHs 的结构重建示意图[35]Fig.3㊀Schematic diagram of CLDHs structural reconstruction [35]1140㊀水泥混凝土硅酸盐通报㊀㊀㊀㊀㊀㊀第42卷对比四种常用的水滑石制备方法可知,共沉淀法操作便捷,制得LDHs 结晶度好,其余方法均需要LDHs 作为前驱体进行制备㊂然而,对LDHs 进行改性通常采用焙烧还原法,这归因于其充分利用LDHs 的热稳定性和记忆效应等性质㊂1.3㊀水滑石复合水泥材料吸附机理图4㊀不同处理方法下Mg-Al-CO 3LDHs 的XRD 谱[37]Fig.4㊀XRD patterns of Mg-Al-CO 3LDHs under different treatment methods [37]LDHs 的层状结构和离子交换等特性,使其具有显著的离子吸附性能㊂其中,LDHs 的阴离子吸附能力较CLDHs 弱,可归因于焙烧处理后的LDHs 活性中心增加,对介质溶液中阴离子和水分的吸附能力增强[36],图4为Mg-Al-CO 3LDHs 在不同处理方法下的物相分析,可观察到原状LDHs(即O-LDHs)表现出良好的结晶状态,在经过600ħ的高温焙烧后,LDHs 层状结构发生改变,LDHs 衍射峰由MgO 所取代㊂而经过再水化过程的焙烧水滑石R-LDHs 的衍射峰型和角度与O-LDHs 基本一致[37]㊂离子交换平衡不仅受LDHs 的类型㊁摩尔比和制备方法等因素的影响,还受介质溶液浓度㊁温度和pH值的影响[38-39]㊂目前,众多学者采取吸附等温线确定平衡吸附量㊂最常用的平衡吸附模型分别为Langmiur 和Freundlich 模型,其中,Langmuir 等温线是均匀吸附表面的单分子吸附模型㊂Zuo 等[40]㊁Xu 等[41-42]㊁Chen 等[43]㊁Yoon 等[44]研究均表明Langmiur 吸附等温线模型与试验数据的拟合效果好,可归因于LDHs 的电中性,每个吸附位点只能固定一个氯离子,等温吸附发生在均匀表面㊂研究结果表明,插层为NO -3的LDHs 氯离子吸附量普遍大于插层为NO -2的LDHs 氯离子吸附量,这是由于NO -2插层的LDHs 基底间距小于NO -3插层的LDHs,这使得NO -2插层的LDHs 进行阴离子交换相对困难[41]㊂此外,在同一介质溶液中进行吸附时,共沉淀法吸附效率高于焙烧还原法和水热法㊂对于焙烧还原法,这可能是焙烧过程中温度对结构的破坏引起的㊂水热法吸附量低的原因在于交换过程受产物中的杂质和环境中CO 2的影响㊂当溶液中存在其他竞争性阴离子(如SO 2-4㊁CO 2-3),且浓度高于氯离子浓度时,竞争性阴离子会干扰氯离子吸附,对LDHs 的吸附氯离子性能产生不利影响[42-43]㊂不同类型LDHs 氯离子吸附量见表1㊂表1㊀不同类型LDHs 氯离子吸附量Table 1㊀Chloride ion adsorption capacity of different types of LDHsLDHs type Molar ratio of M 2+to M 3+Synthetic method Solution composition pH value Experimental condition Adsorption capacity /(mmol㊃g -1)Ref.MgAl-NO 21ʒ1Coprecipitation NaCl 7Room temperature 3.61[40]MgAl-NO 23ʒ1Calcination-rehydration NaCl 7Room temperature 1.6[40]MgAl-NO 21ʒ1Hydrothermal NaCl 7Room temperature 2.55[40]MgAl-NO 22ʒ1Ion exchange Ca(OH)2+NaCl 12.625ħN 2atmosphere 2.51[41]MgAl-NO 32ʒ1Coprecipitation Ca(OH)2+NaCl 12.625ħN 2atmosphere 3.61[41]MgAl-NO 32ʒ1Coprecipitation NaCl +NaOH 1325ħN 2atmosphere 2.74[42]MgAl-NO 32ʒ1Coprecipitation NaCl +NaOH +Na 2SO 41325ħN 2atmosphere 0.65[42]CaAl-NO 32ʒ1Coprecipitation NaCl 7Room temperature 25ħ 3.38[43]CaAl-NO 32ʒ1Coprecipitation Cement mortar 4.5[43]MgAl-pAB 1ʒ1Calcination-rehydration Cement paste Room temperature 20ħ 4.31[44]2㊀不同类型水滑石的氯离子吸附性能2.1㊀焙烧改性水滑石基于LDHs 的阴离子交换㊁记忆效应以及结构重建等特性,可以推断出,与LDHs 相比,经过高温焙烧后的CLDHs 的吸附氯离子性能更优㊂这归因于CLDHs 层间出现较多的活性中心,可吸附较多的氯离子,从而第4期周丽娜等:水滑石复合水泥基材料氯离子吸附能力的研究进展1141㊀完成层状结构的重建[44]㊂图5(a)和(b)分别为焙烧改性原状水滑石CLDHs 和在水泥基材料中完成结构重建的LDHs 的SEM 照片㊂从微观形貌上看,前者样品颗粒边缘尖锐锋利,而后者在水泥基质中完成离子交换后,具有明显的片状结构且层层堆叠在一起㊂图5㊀不同环境中水滑石在水泥基质中的SEM 照片[44]Fig.5㊀SEM images of LDHs in different environment conditions in cement paste[44]图6㊀钢筋在模拟孔隙液及CLDHs 处理前后的Nyquist 图[46]Fig.6㊀Nyquist plots for rebar in simulated pore solution with /without CLDHs [46]混凝土中掺入的CLDHs 吸附氯离子,可提高钢筋表面的氯离子阈值,从而延缓钢筋锈蚀开始的时间㊂图6为钢筋在模拟孔隙液(simulated pore solution,SPS)及CLDHs 处理前后的Nyquist 图,唐聿明等[45]和牛乐[46]发现经CLDHs 处理的孔隙液,电容弧直径下降明显,这表明处理后的孔隙液对钢筋的侵蚀作用减小㊂水滑石在水泥基材料体系与模拟孔溶液中发挥的固氯作用存在不一致的结论㊂CLDHs 对氯离子的吸附是利用其结构记忆效应的化学吸附,胡静等[47]和张琳[48]针对MgAl-CO 3CLDHs,在模拟孔隙液中通过吸附热力学对其吸附氯离子机理进行基础研究㊂在此基础上,张琳进一步探究了CLDHs 在水泥净浆中的固氯能力和机理,结果表明,与对照组相比,掺CLDHs 的试验组固氯能力更好,值得关注的是,CLDHs 固氯能力随养护龄期的增长呈增长趋势,且固氯能力与氯离子浓度呈正相关㊂混凝土中的氯离子可分为结合氯离子和自由氯离子,一般认为自由氯离子会破坏钢筋界面的钝化膜[49]㊂王佩[50]研究表明,掺入的CLDHs 填充水泥基材料的孔隙,其密实度提升,抗冻能力增强,固氯能力随LDHs 掺量的增加而提高㊂基于LDHs 的离子交换功能,段平等[51]㊁Ma 等[52]和Shui 等[37]评估了不同处理方式得到图7㊀掺LDHs 混凝土氯离子扩散系数[58-60]Fig.7㊀Chloride diffusion coefficient of concrete mixed with LDHs [58-60]的CLDHs 混凝土的抗碳化性能㊂上述研究结果表明,CLDHs 的吸附氯离子性能优于LDHs㊂Liu 等[53]验证了这一结论㊂作为替代水泥的外掺材料,LDHs 的掺量须结合水泥基材料拌合物的工作性能进行调节和控制㊂冯跃等[54]㊁Wang 等[55]和宋学锋等[56]分别针对水滑石复合水泥基材料的氯离子侵蚀和碳化问题进行了相应研究㊂在不影响砂浆整体强度的条件下,CLDHs 的最佳掺量控制在胶凝材料质量的1%~3%㊂过量掺入LDHs 会影响水泥水化进程,对水泥基材料的孔隙结构和强度造成不利影响,从而影响水泥基材料抗氯离子渗透的能力[57]㊂图7为学者们采用快速氯离子迁移1142㊀水泥混凝土硅酸盐通报㊀㊀㊀㊀㊀㊀第42卷法表征不同LDHs 掺量下的水泥基材料的氯离子吸附能力[58-60]㊂由图7可知,混凝土中氯离子迁移系数D RCM 随LDHs 的掺量的增加呈先减小后增大的趋势㊂氯离子迁移系数减小的原因在于LDHs 的掺入降低混凝土基体孔隙率,提高其抗氯离子渗透能力,而LDHs 比表面积较大,拌和时吸附水泥基体中的水分,过量掺入使水化反应不充分,Ca(OH)2的生成量减少,对混凝土的孔隙结构产生不利影响,进而导致抗氯离子渗透能力降低㊂图8为段平[36]㊁马军涛[58]㊁Ma 等[61]㊁Wang 等[62]采用NT-Build 443法测定氯盐环境下不同LDHs 掺量混凝土氯离子浓度分布情况㊂由图8可知,复掺LDHs 混凝土内部不同深度氯离子浓度均呈递减趋势,且较对照组浓度低㊂呈这种变化趋势的原因:一方面是LDHs 的掺入会减少混凝土内部的毛细孔数量,改善混凝土的微观结构,阻碍氯离子通过孔隙向混凝土内部扩散的进程;另一方面是LDHs 会吸附部分氯离子,延缓了氯离子向内部渗透的进程㊂图8㊀氯盐环境下掺LDHs 试块的氯离子分布曲线[36,58,61-62]Fig.8㊀Chloride ion distribution curves of specimens mixed with LDHs in chloride salt environment [36,58,61-62]2.2㊀阻锈阴离子插层水滑石阻锈阴离子吸附效能取决于其自身同外界氯离子之间的亲和力顺序[63-64],即LDHs 层间阴离子亲和力越低,越容易被外界氯离子所代替,其固氯能力就越强,对钢筋防腐性能和混凝土耐久性能提升效果就越显著㊂因此,国内外学者通过不同方法合成具有阻锈能力或缓蚀能力的LDHs,探讨其吸附氯离子的能力㊂其中,可发现插层多为NO -3或NO -2㊂从XRD 的角度对插层分别为NO -3和NO -2的LDHs 的氯离子吸附能力进行研究,如图9所示㊂由图可知,二者在氯离子吸附前后,均存在明显的LDHs 衍射特征峰,这表明样品结晶良好㊂从图9(a)可以看出,在氯离子吸附前后,LDHs-NO -3基底间距从0.8717nm 降低至0.7890nm,这是由于氯离子直径小于NO -3的直径;从图9(b)可以看出,在氯离子吸附前后,LDHs-NO -2基底间距从0.7702nm 略升至0.7771nm,这是由于氯离子与NO -2在直径上较相近㊂因此,在进行离子交换时,LDHs-NO -3较LDHs-NO -2相对容易,且前者的氯离子吸附量大于后者㊂Yang 等[60]㊁Cao 等[65]㊁Chen 等[66]和Yang 等[67]制备了插层阴离子具有缓蚀效果的LDHs,均可在钢筋混凝土中发挥双重效应,捕捉水泥基材料中的游离氯离子,同时向水泥基材料中释放阻锈阴离子,有效控制混凝土中钢筋的锈蚀,但是须考虑LDHs 的掺量,Yang 等[67]指出过量使用LDHs 会增加混凝土的孔隙率,为氯离子的传输提供途径㊂ZnAl-NO 2LDHs 缓蚀性能优异,其对钢筋的缓蚀效率受pH 值的影响较显著㊂Gomes 等[68]认为孔隙液在中性pH 值条件下,ZnAl-NO 2LDHs 具有良好的氯捕集能力,但随着pH 值的增加,氯捕集能力降低,这是由于孔隙液中的OH -会占据LDHs 结构中的结合位点㊂Cao 等[69]认为孔隙液在碱性pH 值条件下,插层离子的释放不仅受OH -对LDHs 的亲和力的影响,也受LDHs 部分溶解的影响,但目前,关于LDHs 在碱性溶液中发生部分溶解对缓蚀率的影响还未开展系统研究㊂此外,田玉琬等[70]认为可根据孔隙液中氯离子浓度智能释放NO -2,高碱环境下的释放效果主要受氯离子控制㊂第4期周丽娜等:水滑石复合水泥基材料氯离子吸附能力的研究进展1143㊀图9㊀氯离子吸附前后LDHs 的XRD 谱[41]Fig.9㊀XRD patterns of LDHs before and after chloride adsorption [41]2.3㊀复合防御体系水滑石LDHs 在水泥基材料中的形成机制包括两种:受掺合料中镁铝氧化物相对含量影响的MgAl-LDHs 和具有水化产物AFm 相结构的CaAl-LDHs [71]㊂Kayali 等[72]通过快速氯离子渗透法㊁XRD 等测试方法对掺有矿渣的混凝土试样进行研究,发现LDHs 作为一种重要的水化产物在矿渣混凝土中生成,并显示出优越的固氯能力,以延缓混凝土结构中的钢筋锈蚀㊂LDHs 吸附侵蚀阴离子,复合掺入矿物掺合料,发挥二者的叠加效应,使胶凝材料的颗粒级配更为合理,在微集料效应作用下,完善混凝土内部微观结构,显著提升混凝土综合性能㊂部分学者通过矿物掺合料与LDHs 复掺提高混凝土的抗氯离子侵蚀能力及耐久性,并取得了一定的成果㊂图10㊀复合体系在氯盐环境下的氯离子分布曲线[58,73]Fig.10㊀Chloride ion distribution curves of composite system in chloride salt environment [58,73]混凝土抗氯离子侵蚀性能测试可按照NTBuild-443非稳态自然扩散法进行,马军涛[58]和陈宇轩等[73]采用此法研究了LDHs 与偏高岭土(metakaolin,MK)为主的矿物掺合料复合使用,对混凝土抗氯离子渗透性能的改善情况,并得出氯离子扩散浓度同扩散距离二者之间的规律㊂图10为二者采用NTBuild-443的数据对比,由图可见,相较于对照组,复合体系下的LDHs提升了混凝土抗氯离子侵蚀的能力㊂LDHs 材料主要通过煅烧处理后的结构重建过程和离子交换实现对氯离子的吸附,掺加改性剂的混凝土可进一步改善其抗氯离子渗透性,提高混凝土的抗氯离子侵蚀能力㊂陈国玮等[74]发现LDHs 材料可有效吸附外界侵入的SO 2-4,而MK 在混凝土中的火山灰特性可促进水泥的水化过程,改善混凝土内部的显微结构㊂3㊀结语与展望本文从LDHs 结构特性角度出发,综述了不同类型LDHs 体系对氯离子吸附能力的研究现状,对于探明复掺LDHs 水泥基材料体系的固氯能力和延缓钢筋锈蚀的机理提供了理论基础,得到的主要结论如下:1)LDHs 的氯离子吸附能力受制备工艺的影响,通过吸附等温模型评估了不同制备方法所得的LDHs 氯离子吸附量㊂其中,Langmiur 模型对试验数据的拟合效果最好,且共沉淀法制备的LDHs 氯离子吸附量优于水热法和焙烧还原法㊂2)水泥基材料孔隙液的pH 值和氯离子浓度对LDHs 的氯离子吸附性能有一定的影响㊂当孔隙液在高碱环境下或存在其他竞争性离子,且浓度高于氯离子浓度时,竞争性离子会率先占据水滑石层间的结合位1144㊀水泥混凝土硅酸盐通报㊀㊀㊀㊀㊀㊀第42卷点,影响LDHs对孔隙液中氯离子的捕捉㊂3)LDHs的氯离子吸附能力主要取决于其独特的层状结构和离子交换性质㊂CLDHs利用记忆效应,在结构重建的过程中完成对水泥基材料孔隙液中氯离子吸附,阻锈阴离子插层LDHs利用层间离子交换功能,在捕捉氯离子的同时,释放具有缓蚀效果的阴离子,高温焙烧处理的水滑石对氯离子吸附效果更好;复合防御体系LDHs发挥叠加效应,在LDHs吸附机理基础上,从改善孔隙结构的角度改善氯离子侵蚀㊂然而,由于实际试验的局限,对不同种类的LDHs的吸附效果难以进行精确的量化分析㊂目前,关于LDHs在碱性孔隙液中发生部分溶解对缓蚀离子的释放速率影响,以及不同矿物掺合料与LDHs复合对水泥基材料的氯离子吸附效应和机理还需进一步系统研究㊂参考文献[1]㊀杨长辉,晏㊀宇,欧忠文.偏高岭土水泥净浆结合氯离子性能的研究[J].混凝土,2010(10):1-3+7.YANG C H,YAN Y,OU Z W.Capability of cement paste binding chloride ions with metakaolin as admixture[J].Concrete,2010(10):1-3+ 7(in 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dft计算在燃料电池中的应用英文回答：DFT (Density Functional Theory) is a powerful computational method used in various fields of science and engineering, including the study of fuel cells. Fuel cells are electrochemical devices that convert the chemical energy of a fuel, such as hydrogen, into electrical energy. Understanding the behavior of fuel cell materials at the atomic level is crucial for improving their efficiency and performance.One of the key applications of DFT in fuel cell research is the prediction and optimization of catalyst materials. Catalysts play a critical role in fuel cell reactions by facilitating the electrochemical reactionsthat occur at the electrodes. DFT calculations can be used to investigate the electronic structure and reactivity of different catalyst materials, providing insights into their catalytic activity and selectivity. By analyzing the energyprofiles of the reaction pathways, researchers can identify the most promising catalysts for specific fuel cell reactions.Another important application of DFT in fuel cell research is the study of fuel cell membranes. Membranes are essential components of fuel cells as they separate thefuel and oxidant streams while allowing the transport ofions necessary for the electrochemical reactions. DFT calculations can be used to understand the transport properties of different membrane materials, such as proton conductivity and oxygen permeability. This information can guide the development of new membrane materials with improved performance and durability.Furthermore, DFT can also be used to investigate the interactions between fuel cell materials and impurities or contaminants. For example, DFT calculations can be employed to study the adsorption of carbon monoxide (CO) on the catalyst surface, a common impurity in fuel cell feedstocks. By understanding the adsorption behavior of CO, researchers can design catalyst materials that are more resistant topoisoning and improve the overall stability and longevity of the fuel cell system.In summary, DFT calculations have a wide range of applications in fuel cell research, including catalyst design, membrane optimization, and understanding material interactions. By providing atomic-level insights into the properties and behavior of fuel cell materials, DFT can contribute to the development of more efficient and durable fuel cell systems.中文回答：DFT（密度泛函理论）是一种在科学和工程的各个领域中广泛应用的强大计算方法，包括燃料电池的研究。

胶体对砷的吸附作用及其影响因素研究现状与进展（黄臣臣2016021377）摘要：砷因其具有极强的毒性和致癌作用引起全世界的关注。

土壤胶体由于其独特的性质，是土壤环境中重要的污染物吸附和运输载体，对砷的分配和运移起到关键的作用。

因此，本文主要从土壤胶体的基本性质；不同形态As的吸附；影响As吸附的因素，如pH、Eh、土壤有机碳，重点分析共存离子对As的吸附作用；As吸附研究中应用的技术和模型，如傅里叶红外光谱、表面络合模型等几个方面，对共存离子对砷在土壤胶体上吸附的影响进行论述。

由于实验控制条件的差异，各个因素对砷在胶体上的吸附影响研究结果不同，大部分学者认为磷酸根、硅酸根、硝酸根等共存离子会与磷酸产生竞争作用而降低其吸附量，但仍有研究表明，共存离子对胶体结合态砷没有影响，由此可见，要清楚其中缘由，还需要结合室内室外、微观及宏观实验进一步深入研究。

关键词：土壤胶体砷共存离子竞争吸附红外光谱Abstract：As is of concern byall over the world due to its highly toxic and carcinogenic . Soil colloids is an important pollutants adsorption and carrier in soil environmentbecause of its unique characteristics, which play a key role in the arsenic distribution and migration. Therefore, this article mainly discuss the basic properties of soil colloid, adsorption of different forms of As. The factors that influence the adsorption of As, such as pH, Eh, soil organic carbon, mainly focuses on the analysis of coexisting ions on As adsorption. Moreover, topics for technology and model application in the study of arsenic adsorption are also given, such as Fourier transform infrared spectroscopy and surface complexation. The essay tries to expound the effects of concomitant ions on the As adsorption process in soils colloid from the above-mentioned several aspects. Due to the differences of the experimental conditions, the influence of various factors on the adsorption of arsenic on colloids is different. Most scholars believe that the phosphate, silicate, nitrate and phosphate coexistions will have competitive effects and reduce its adsorption capacity, but coexisting ions have been found to be limited or insignificant. Thus, it is necessary to do some further research combine indoor and outdoor, micro and macro experiments tounderstand the reasons.Keywords: soil colloids, arsenic, coexisting ions, competitive adsorption, infrared spectroscopy1.引言砷具有极强的毒性和致癌作用，无机As(Ⅲ)和As(Ⅴ)被认为是一级致癌物质。

第49卷第2期2021年1月广㊀州㊀化㊀工Guangzhou Chemical IndustryVol.49No.2Jan.2021木质材料对铜离子吸附的研究进展∗周㊀彤,梁建军,王㊀磊,刘义章,汪蓓蓓(滁州职业技术学院,安徽㊀滁州㊀239000)摘㊀要:木质材料主要成分为木质素㊁纤维素等物质,其表面含有大量的羟基㊁羰基等活性基团,这些活性基团含有可以与Cu 2+空轨道配位的孤对电子对㊂通过制备活性炭和化学改性的方式能增加木质材料的比表面积和吸附位点,因而可显著提高吸附剂对Cu 2+的吸附性能㊂同时这些木质材料凭借来源广泛㊁价格低廉㊁可再生㊁无污染等优势成为吸附法的最佳原材料㊂本文综述了国内天然木质材料㊁改性木质材料在去除Cu 2+方面的研究进展,为木质材料的应用研究提供参考㊂关键词:吸附;锯末;铜离子;改性㊀中图分类号:X703.1㊀㊀㊀㊀㊀文献标志码:A㊀㊀㊀㊀文章编号:1001-9677(2021)02-0021-03㊀㊀㊀㊀㊀㊀㊀㊀㊀㊀㊀㊀㊀∗基金项目:滁州职业技术学院2019年校级科研一般项目(YJY -2019-11);滁州职业技术学院2020年校级科研重点项目(YJZ -2020-03)㊂第一作者:周彤(1990-),女,硕士研究生,助教,从事环境水污染处理研究㊂Research Progress on Adsorption of Cu 2+by Wood Materials ∗ZHOU Tong ,LIANG Jian -jun ,WANG Lei ,LIU Yi -zhang ,WANG Bei -bei (Chuzhou Vocational and Technical College,Anhui Chuzhou 239000,China)Abstract :The main components of wood materials are lignin,cellulose and other substances,and its surface contains a large number of active groups such as hydroxyl and carbonyl groups.These active groups contain lone pairs of electrons that can coordinate with the empty Cu 2+orbitals.The preparation of activated carbon and chemical modification can increase the specific surface area and adsorption sites of wood materials,thus it can significantly improve the adsorption performance of the adsorbent for Cu 2+.At the same time,these wooden materials have become the best raw materials for the adsorption by virtue of their wide -ranging sources,low price,renewable,and pollution -free advantages.The research progress on application of natural wood materials,and its modification materials on absorption of Cu 2+were summarized,in order to provide references for the application research of wood materials.Key words :absorption;sawdust;copper ions;modified正所谓,水是生命的源泉㊂人类的一切生产活动都离开水㊂我国是一个人口大国,同时也是个淡水资源匮乏的国家㊂在淡水资源匮乏的同时,我国的水污染情况也不容乐观㊂水污染不仅造成了数额巨大的经济损失,更是直接危害了人们的饮用水安全[1]㊂水俣病和痛痛病让人类第一次认识到重金属污染的危害,也让学者将注意力吸引到重金属污染的处理上㊂重金属是指密度大于5kg /m 3以上的金属[2],大约有45种㊂我们通常所说的重金属是从环境污染角度分类的,指的是具有生物毒性的金属如汞㊁镉㊁铅㊁铬㊁类金属砷等[3]㊂当少量含Cu 2+的污水进入水体,由于水体有一定的自净能力,对水体中生物影响不大㊂但当能产生重金属Cu 2+的行业如采矿业㊁电镀㊁造纸㊁电池㊁化肥㊁皮革㊁农药等将大量的含Cu 2+废水排入自然界中,就可能造成水体污染㊂重金属离子具有生物不可降解性,并能在生物体内蓄积,因而危害生物健康㊂目前,水体中Cu 2+的除去方法主要有化学沉淀法㊁离子交换法㊁电化学法㊁吸附法[4]等㊂化学沉淀法和电化学法在使用时,需要较高浓度的Cu 2+,同时产生的化学污泥还可能存在二次污染等问题㊂吸附法因对污染物的浓度要求不高,方法简单㊁容易操作等特点,在重金属去除方面越来越受到重视㊂同时,还可以利用化学改性的方法,将特定的官能团引入吸附剂,来提高吸附性能或者选择性地去除某一类污染物㊂林业生产为人类社会提供优质木材的同时也产生了大量木材剩余物如锯末㊁木屑㊁枝条㊁树皮等㊂据报道木材的利用率仅有10%㊂如何利用好剩余的90%是全人类需要解决的难题㊂这些废弃物可于生产工业酒精㊁活性炭,制备燃烧棒,提取化肥等㊂此外,这些废弃物凭借来源广泛㊁价格低廉㊁无污染等优势,还可用做吸附法的原材料㊂锯末和木屑是典型的林业生产的废弃物,其主要成分为木质素㊁纤维素等物质,故其表面含有大量的羟基㊁羰基等活性基团,是一种应用前景广泛的重金属吸附剂附[5-7]㊂这些木质材料既可直接用作吸附剂原料也可通过化学改性㊁制备活性炭等方式来提高它的吸附性能㊂研究木质材料吸附机理有利于科研工作者掌握其吸附的本质,在改性时针对目标污染物有目的性地增加木质材料的表面特征官能团,进而提高吸附性能㊂天然木质材料其表面含有大量的羟基㊁羰基等活性基团,这些基团可以与重金属Cu 2+配22㊀广㊀州㊀化㊀工2021年1月位㊂如果在配位的过程中有其他离子析出时,那么天然木质材料的吸附机理除配位作用外,还存在离子交换㊂木质材料活性炭相比木质材料表面更粗糙㊁孔径大㊁比表面积大,故木质材料活性炭吸附Cu2+的机理主要为物理吸附,此时的吸附作用力主要为范德华力㊂化学改性木质材料的吸附机理因改性方法的不同而不同㊂氨基改性木质材料的吸附机理主要是化学吸附,即氨基上N原子和Cu2+的配位作用㊂酸改性天然木质材料可以增加其表面羟基的含量,利用配位作用增加吸附性能㊂碱改性天然木质材料经过碱处理后经过一系列物理化学变化后,木质材料结构疏松㊁孔径变大,同时碱可以提供㊃OH,故在物理吸附和化学吸附的双重作用下,对Cu2+吸附性能增加显著㊂无论是哪种改性方法,其吸附机理都不是单一的,只是改性方法的不同使得某一种机理为主导㊂这也从另一方面说明天然木质材料的吸附较某种单一组分的吸附剂吸附过程更加复杂㊂1㊀利用天然木质材料特定组成去除水体中的Cu2+㊀㊀锯末的骨架结构成分主要有纤维素㊁木质素㊁半纤维素等,内部表面富含羟基㊁羰基㊁甲氧基㊁羧基等基团[8]㊂木屑中含有大量的官能团如羟基㊁羧基㊁氨基酸和木质酚,它们都能吸附部分阳离子[9]㊂这些木质材料因木材的种类不同,其主要成分及其含量有所差异,对Cu2+的吸附性能也有所不同㊂天然的木质对Cu2+有一定的吸附能力主要是材料表面含有羟基㊁羰基等活性基团,这些基团含有可以与Cu2+空轨道配位的孤对电子㊂刘晓凤等[10]利用海南椰子树木屑为原料制备出粒径大小不同的木屑粉末并研究了其对铅㊁铜㊁镉三种重金属离子的吸附行为,结果表明木屑吸附铜离子的吸附量比较小,饱和吸附量都没超过5mg/g㊂郜洪文[11]利用成材较快的竹子作为原材料来除去废水中的污染物Cu2+,研究表明竹锯末对Cu2+的吸附速度很快,吸附机制主要为离子交换,该研究为竹锯末综合利用提供新途径㊂孙杰等[12]用原松树锯末和柠檬酸钠改性锯末对重金属离子Cu2+的吸附行为进行研究,结果表明在最佳条件下原松树锯末㊁柠檬酸钠改性锯末对Cu2+的最大吸附量分别为3.31mg/g㊁5.75mg/g㊂由上可知,天然木质材料对Cu2+的吸附量较低,吸附量也就几个毫克每克,这在一定程度上限制了天然木质材料的相关研究应用㊂但是作为一种低廉㊁易得㊁可再生的农林废弃物,木然木质材料仍是一种潜在的吸附剂材料㊂2㊀利用比表面或空隙作用来去除水体中的Cu2+活性炭因比表面积大㊁多孔而具有良好的吸附性能,是目前商业应用最为广泛的吸附剂㊂木质活性炭是木质材料经过相关化学处理后,利用高温将其碳化,碳化后的木质材料表面粗糙㊁内部多孔㊁比表面积变大㊂根据活性炭制备过程中所使用原料的不同,可将其分为生物质活性炭㊁木质活性炭㊁合成材料活性炭以及煤质活性炭[13]㊂木屑㊁锯末因价格低廉㊁来源广泛㊁可再生㊁无污染是制作木质活性炭的绝佳原材料㊂限氧升温碳化法㊁热解法是制备木质材料活性炭的主要方法㊂黄宏霞等[14]以木屑为原料㊁磷酸为活化剂㊁硼酸为催化剂制备出化学改性木屑活性炭,并将此活性炭用于Cu2+溶液的吸附研究,活性炭的理论最大吸附量高达14225mg/kg㊂周丹丹等[15]利用松木屑为原料采用限氧升温碳化法,分别在200㊁300㊁400㊁500ħ四个温度进行碳化制备出4种生物炭,并将4种生物炭用于Cu2+的吸附研究㊂研究表明松木生物炭在热解温度为200ħ时对Cu2+的吸附性能最好㊂生物质热解是指在惰性气体保护的氛围下,对生物质进行加热,当达到生物质热解温度时,生物质分解得到气态挥发分和固态炭[16],通过冷凝的方式可以将气态挥发物变成焦油㊂目前在生物质热解主要采用直接热解和化学物质浸泡后再热解两种方式㊂毛明翠等[17]利用苹果树枝和梧桐木锯末为原料,采用450ħ热裂解法制备出生物炭,并用于铜离子的吸附研究㊂NaOH是最常用的化学物质浸泡液之一㊂蔡静[18]采用热解+ NaOH的方法对松木屑进行改性并探求了改性热解炭对Cu2+的吸附,结果表明改性热解炭在吸附时间和吸附性能都优于原木屑和直接热解炭㊂3㊀利用化学改性方法提高木质材料特定组成来除去水体中的Cu2+3.1㊀氨基改性木质材料将氨基引入木质材料的表面,即把对重金属离子能产生强配位作用的N原子引入木质材料的表面,可以利用N原子和Cu2+之间的强配位作用来提高木质材料的吸附性能㊂此外,在酸性条件下,氨基对一些阴离子染料也有很好的吸附性能㊂目前使用较多的氨基改性材料有二乙烯三胺㊁聚乙烯亚胺㊁乙二胺等㊂学者多利用环氧氯丙烷交联或直接反应等方式将氨基引入木质材料㊂刘梦珠等[19]对杨木屑用NaClO预处理并用二乙烯三胺进行氨基化改性制备出改性木屑,并研究了其对水中Cu(Ⅱ)/Cr(Ⅵ)的连续吸附研究㊂结果表明改性木屑对Cu(Ⅱ)的饱和吸附容量能达到195.70mg/g且吸附过程符合Langmuir模型,较原木屑吸附容量提高显著㊂王平等[20]直接利用木屑与三乙胺的反应来制备改性木屑,同时研究了其对废水中铜㊁镉等重金属离子的吸附情况㊂研究结果表明:吸附条件为温度为30ħ㊁pH为8㊁铜离子浓度小于等于400mg/kg时,改性木屑对铜离子去除率基本保持在99.99%㊂夏璐等[21]使用乙二胺改性的木屑黄原酸盐对水溶液中的Cu(II)㊁Ni(II)离子进行吸附研究,这里采用环氧氯丙烷交联的方式将氨基引入木屑中㊂该吸附过程为单层吸附且吸附过程可以用Langmuir模型和准二级动力学模型进行描述,计算得到吸附过程的活化能59.12kJ/mol,这些参数都表明该吸附过程为化学吸附㊂由上可知氨基改性木质材料较原木质材料对Cu2+的吸附性能提高显著,且吸附机理主要为化学吸附㊂这也与N原子和Cu2+之间的强配位作用相一致㊂3.2㊀酸改性木质材料酸改性木质材料主要是利用硫酸㊁磷酸㊁硝酸㊁醋酸等常见酸,采用浸泡㊁搅拌等方式来改变木质材料表面物质组成,进而提高木质材料的吸附性能㊂同时酸改性木质材料具有可燃性好的优势,也有利于采用热法回收木质材料上吸附的重金属[22]㊂杜玉辉等[23]在40ħ时将锯末与已知浓度的硝酸进行反应制备出酸改性锯末,并研究了锯末用量㊁Cu2+浓度㊁溶液pH㊁吸附时间等因素对吸附过程的影响㊂3.3㊀碱改性木质材料木质材料的主要成分纤维素可以与碱金属氢氧化物溶液发第49卷第2期周彤,等:木质材料对铜离子吸附的研究进展23㊀生一系列的物理化学变化,使得纤维素的润胀[24],比表面积增加显著㊂王卓然等[25]利用KOH溶液对松木屑进行碱改性制备出改性木屑,改性木屑对Cu2+的吸附效率较未改性前由72%提高到97%,吸附性能提高明显㊂阳康[26]用NaOH㊁异丙醇组合剂浸泡锯末制备出碱改性锯末,与天然锯末相比碱改性锯末受投量变化的影响较小㊂4㊀结㊀语天然木质材料成分主要有纤维素㊁木质素㊁半纤维素等,其表面和内部含有活性官能团羟基㊁羧基等,对水体中污染物Cu2+有一定的去除能力㊂学者通过制备木质材料活性炭来增大木质材料的比表面积,或通过化学改性的方式来增加木质材料表面的活性官能团进而增加吸附位点㊂木质活性炭制备主要介绍了限氧升温碳化法和热解法两种,其中木质材料热解后用化学物质浸泡吸附性能更好㊂酸改性㊁碱改性以及氨基改性是目前木质材料化学改性采用的主要方法㊂我国是一个农业大国,研究以天然木质材料或改性木质材料对水体中污染物的吸附行为,对我国农林废弃物的处理及循环利用具有积极的意义㊂农林废弃物对水体中污染物吸附性能的探讨是目前多个实验室共同研究的课题,但如何将此研究进行工业化大生产并保持一致的吸附性能仍是每个学者的研究方向和目标㊂水体中重金属污染物往往是多种重金属共存的场景,如何在提高木质材料对其的去除能力和增加选择性地回收某种金属离子方面仍有较大的研究空间㊂参考文献[1]㊀姚诚.水污染现状及其治理措施[J].污染防治技术,2009(2):87-90,96.[2]㊀武文会.腐殖酸对活性污泥吸附铜离子的影响研究[D].重庆:重庆大学,2015.[3]㊀林雪原,荆延德,巩晨,等.生物炭吸附重金属的研究进展[J].环境污染与防治,2014(5):83-87.[4]㊀王月月,李娟英,鲁玉渭,等.响应面优化玉米芯对Cu2+的吸附[J].上海海洋大学学报,2020(3):355-363.[5]㊀常兴涛,岳建芝,贾洋洋,等.锯末颗粒吸附去除低质量浓度氨氮废水的研究[J].河南农业大学学报,2018(4):582-586. 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Adsorption char acter istics of copper , lead, zinc and cadmium ions by tourmaline(环境科学学报英文版) 电气石对铜、铅、锌、镉离子的吸附特性JIANG Kan1,*, SUN Tie-heng1,2 , SUN Li-na2, LI Hai-bo2(1. School of Municipal and Environmental Engineering, Harbin Institute of Technology, Harbin 150090, China. jiangkan522@; 2. Key Laboratory of Environmental Engineering of Shenyang University, Shenyang 110041, China)摘要：本文研究了电气石对Cu2+、Pb2+、Zn2+和Cd2+的吸附特性，建立了吸附平衡方程。

研究四种金属离子的吸附等温线以及朗缪尔方程。

结果表明电气石能有效地去除水溶液中的重金属且具有选择性：Pb2+> Cu2+> Cd2+> Zn2+。

电气石对金属离子吸附量随着介质中金属离子的初始浓度的增加而增加。

电气石也可以增加金属溶液的pH值；发现电气石对Cu2+、Pb2+、Zn2+和Cd2+的最大吸附量为78.86、154.08、67.25和66.67mg/g；温度在25-55℃对电气石的吸附量影响很小。

此外研究了Cu2+、Pb2+、Zn2+和Cd2+的竞争吸附。

同时观察到电气石对单一金属离子的吸附能力为Pb>Cu>Zn>Cd，在两种金属系统中抑制支配地位是Pb>Cu，Pb>Zn，Pb>Cd，Cu>Zn，Cu>Cd，和Cd>Zn。

关键字：吸附；重金属含量；朗缪尔等温线；电气石介绍重金属是来自不同行业排出的废水，如电镀，金属表面处理，纺织，蓄电池，矿山，陶瓷，玻璃。

Journal of Hazardous Materials B137(2006)384–395Removal of copper(II)and lead(II)from aqueoussolution by manganese oxide coated sand I.Characterization and kinetic studyRunping Han a ,∗,Weihua Zou a ,Zongpei Zhang a ,Jie Shi a ,Jiujun Yang baDepartment of Chemistry,Zhengzhou University,No.75of Daxue North Road,Zhengzhou 450052,PR ChinabCollege of Material Science and Engineering,Zhengzhou University,No.75of Daxue North Road,Zhengzhou 450052,PR ChinaReceived 8November 2005;received in revised form 25December 2005;accepted 13February 2006Available online 28February 2006AbstractThe preparation,characterization,and sorption properties for Cu(II)and Pb(II)of manganese oxide coated sand (MOCS)were investigated.A scanning electron microscope (SEM),X-ray diffraction spectrum (XRD)and BET analyses were used to observe the surface properties of the coated layer.An energy dispersive analysis of X-ray (EDAX)and X-ray photoelectron spectroscopy (XPS)were used for characterizing metal adsorption sites on the surface of MOCS.The quantity of manganese on MOCS was determined by means of acid digestion analysis.The adsorption experiments were carried out as a function of solution pH,adsorbent dose,ionic strength,contact time and temperature.Binding of Cu(II)and Pb(II)ions with MOCS was highly pH dependent with an increase in the extent of adsorption with the pH of the media inves-tigated.After the Cu(II)and Pb(II)adsorption by MOCS,the pH in solution was decreased.Cu(II)and Pb(II)uptake were found to increase with the temperature.Further,the removal efficiency of Cu(II)and Pb(II)increased with increasing adsorbent dose and decreased with ionic strength.The pseudo-first-order kinetic model,pseudo-second-order kinetic model,intraparticle diffusion model and Elovich equation model were used to describe the kinetic data and the data constants were evaluated.The pseudo-second-order model was the best choice among all the kinetic models to describe the adsorption behavior of Cu(II)and Pb(II)onto MOCS,suggesting that the adsorption mechanism might be a chemisorption process.The activation energy of adsorption (E a )was determined as Cu(II)4.98kJ mol −1and Pb(II)2.10kJ mol −1,respectively.The low value of E a shows that Cu(II)and Pb(II)adsorption process by MOCS may involve a non-activated chemical adsorption and a physical sorption.©2006Elsevier B.V .All rights reserved.Keywords:Manganese oxide coated sand (MOCS);Cu(II);Pb(II);Adsorption kinetic1.IntroductionThe presence ofheavy metals in the aquatic environment is a major concern due to their extreme toxicity towards aquatic life,human beings,and the environment.Heavy metal ions from wastewaters are commonly removed by chemical precipitation,ion-exchange,reverse osmosis processes,and adsorption by activated carbon.Over the last few decades,adsorption has gained importance as an effective purification and separation technique used in wastewater treatment,and the removal of heavy metals from metal-laden tap or wastewater∗Corresponding author.Tel.:+8637167763707;fax:+8637167763220.E-mail address:rphan67@ (R.Han).is considered an important application of adsorption processes using a suitable adsorbent [1,2].In recent years,many researchers have applied metal oxides to adsorption of heavy metals from metal-laden tap or wastewa-ter [3].Adsorption can remove metals over a wider pH range and lower concentrations than precipitation [4].Iron,aluminum,and manganese oxides are typically thought to be the most important scavengers of heavy metals in aqueous solution or wastewater due to their relatively high surface area,microporous structure,and possess OH functional groups capable of reacting with met-als,phosphate and other specifically sorbing ions [5].However,most metal oxides are available only as fine powders or are gener-ated in aqueous suspension as hydroxide floc or gel.Under such conditions,solid/liquid separation is fairly difficult.In addition,metal oxides along are not suitable as a filter medium because of0304-3894/$–see front matter ©2006Elsevier B.V .All rights reserved.doi:10.1016/j.jhazmat.2006.02.021R.Han et al./Journal of Hazardous Materials B137(2006)384–395385their low hydraulic conductivity.Recently,several researchers have developed techniques for coating metal oxides onto the surface of sand to overcome the problem of using metal oxides powers in water treatment.Many reports have shown the impor-tance of these surface coatings in controlling metal distribution in soils and sediments[3,6,7].In recent years,coated minerals have been studied because of their potential application as effective sorbents[3,8,9].Iron oxide coated meterials for heavy metal removal have been proved successful for the enhancement of treatment capacity and efficiency when compared with uncoatedfilter media,such as sil-ica sand[10–14],granular activated carbon[15]and polymeric media[16,17].For example,Edwards and Benjamin[7]found that coated media have similar properties to unattached coating materials in removing metals over a wide pH range,and that Fe oxide coated sand was more effective than uncoated sand.Bai-ley et al.[18]used iron oxide coated sand to remove hexavalent chromium from a synthetic waste stream.The influent contained 20mg l−1Cr(VI)and better than99%removal was achieved. Satpathi and Chaudhuri[19]and Viraraghavan et al.[20]have recently reported on the ability of this medium to adsorb metals from electroplating rinse waters and arsenic from drinking water sources,respectively.Green-Pedersen and Pind reported that a ferrihydrite-coated montmorillonite surface had a larger specific surface area and an increased sorption capacity for Ni(II)com-pared to the pure systems[21].Meng and Letterman[22]found that the adsorption properties of oxide mixtures are determined by the relative amount of the components.They also found that ion adsorption on aluminum oxide-coated silica was better mod-eled assuming uniform coverage of the oxide rather than using two distinct surfaces[23].Lo and Chen[8]determined the effect of Al oxide mineralogy,amount of oxide coating,and acid-and alkali-resistance on the removal of selenium from water.Bran-dao and Galembecket reported that the impregnation of cellulose acetates with manganese dioxide resulted in high removal effi-cient of Cu(II),Pb(II),and Zn(II)from aqueous solutions[24]. Al-Degs and Khraisheh[25]also reported that diatomite and manganese oxide modified diatomite are effective adsorbents for removing Pb2+,Cu2+,and Cd2+ions.The sorption capac-ity of Mn-diatomite was considerably increased compared to the original material for removing the studied metals.Filtration quality of diatomite is significantly increased after modification with Mn-oxides.Merkle et al.[26–28]reported that manganese dioxide coated sand was effective for removal of arsenic from ground water in column experiments.Merkle et al.developed a manganese oxide coating method on anthracite to improve the removal of Mn2+from drinking water and hazardous waste effluent.They generated afilter media with an increased surface area after coating with manganese oxides and found manganese oxide coated media have the ability to adsorb and coprecipi-tate a variety of inorganic species.Stahl and James[29]found their manganese oxide coated sands generated a larger surface area and increased adsorption capability with increasing pH as compared to uncoated silica sand.Additional researchers have been investigated to evalu-ate coating characteristics.X-ray diffraction(XRD),X-ray photoelectron spectroscopy(XPS),Fourier transform infrared spectroscopy(FTIR),transmission electron microscopy(TEM), and scanning electron microscopy(SEM)have been used as well to identify components,distribution,and structure of surface oxide coating[7,9,30,31].An energy dispersive X-ray (EDAX)technique of analysis has been used to characterize metal adsorption sites on the sorbent surface.Typically,oxide was non-uniform over the surface as both the oxide and substratum had been observed[7].The research described here was designed to investigate characteristics of manganese oxide coated sand(MOCS)and test the properties of MOCS as an adsorbent for removing copper(II)and lead(II)from synthetic solutions in batch system.SEM/EDAX,XRD,XPS and BET analysis were employed to examine the properties of adsorption reactions for Cu(II)and Pb(II)ions on MOCS in water.The system variables studied include pH,MOCS dose,ionic strength,contact time and temperature.The kinetic parameters,such as E a,k1,k2, have been calculated to determine rate constants and adsorption mechanism.1.1.Kinetic parameters of adsorptionThe models of adsorption kinetics were correlated with the solution uptake rate,hence these models are important in water treatment process design.In order to analyze the adsorption kinetics of MOCS,four kinetic models including the pseudo-first-order equation[32],the pseudo-second-order equation[33], Elovich equation[34],and intraparticle diffusion model[35] were applied to experimental data obtained from batch metal removal experiments.A pseudo-first-order kinetic model of Lagergen is given as log(q e−q t)=log q e−K1t2.303(1)A pseudo-second-order kinetic model istq t=1(K2q2e)+tq e(2) andh=K2q2e(3) an intraparticle diffusion model isq t=K t t1/2+C(4) and an Elovich equation model is shown asq t=ln(αβ)β+ln tβ(5) where q e and q t are the amount of solute adsorbed per unit adsorbent at equilibrium and any time,respectively(mmol g−1), k1the pseudo-first-order rate constant for the adsorption process (min−1),k2the rate constant of pseudo-second-order adsorption (g mmol−1min−1),K t the intraparticle diffusion rate constant (mmol g−1min−1),h the initial sorption rate of pseudo-second-order adsorption(mmol g−1min−1),C the intercept,αthe initial sorption rate of Elovich equation(mmol g−1min−1),and386R.Han et al./Journal of Hazardous Materials B137(2006)384–395the parameter βis related to the extent of surface coverage and activation energy for chemisorption (g mmol −1).A straight line of log(q e −q t )versus t ,t /q t versus t ,q t versus ln t ,or q t versus t 1/2suggests the applicability of this kinetic model and kinetic parameters can be determined from the slope and intercept of the plot.1.2.Determination of thermodynamic parametersThe activation energy for metal ions adsorption was calcu-lated by the Arrhenius equation [36]:k =k 0exp −E aRT (6)where k 0is the temperature independent factor ing mmol −1min −1,E a the activation energy of the reaction of adsorption in kJ mol −1,R the gas constant (8.314J mol −1K −1)and T is the adsorption absolute temperature (K).The linear form is:ln k =−E aRT+ln k 0(7)when ln k is plotted versus 1/T ,a straight line with slope –E a /R is obtained.2.Materials and methods 2.1.AdsorbentThe quartz sand was provided from Zhengzhou’s Company of tap water in China.The diameter of the sand was ranged in size from 0.99to 0.67mm.The sand was soaked in 0.1mol l −1hydrochloric acid solution for 24h,rinsed with distilled water and dried at 373K in the oven in preparation for surface coating.Manganese oxide coated sand was accomplished by utilizing a reductive procedure modified to precipitate colloids of man-ganese oxide on the media surface.A boiling solution containing potassium permanganate was poured over dried sand placed in a beaker,and hydrochloric acid (37.5%,W HCl /W H 2O )solution was added dropwise into the solution.After stirring for 1h,the media was filtered,washed to pH 7.0using distilled water,dried at room temperature,and stored in polypropylene bottle for future use.2.2.Metal solutionsAll chemicals and reagents used for experiments and anal-yses were of analytical grade.Stock solutions of 2000mg l −1Pb(II)and Cu(II)were prepared from Cu(NO 3)2and Pb(NO 3)2in distilled,deionized water containing a few drops of concen-trated HNO 3to prevent the precipitation of Cu(II)and Pb(II)by hydrolysis.The initial pH of the working solution was adjusted by addition of HNO 3or NaOH solution.2.3.Mineral identificationThe mineralogy of the sample was characterized by X-ray diffraction (XRD)(Tokyo Shibaura Model ADG-01E).Pho-tomicrography of the exterior surface of uncoated sand and man-ganese oxide coated sand was obtained by SEM (JEOL6335F-SEM,Japan).The distribution of elemental concentrations for the solid sample can be analyzed using the mapping analysis of SEM/EDAX (JEOL SEM (JSM-6301)/OXFORD EDX,Japan).The existence of Cu(II)and Pb(II)ions on the surface of manganese oxide coated sand was also confirmed by using EDAX.Samples for EDAX analysis were coated with thin carbon film in order to avoid the influence of any charge effect during the SEM operation.The samples of MOCS and MOCS adsorbed with copper/lead ions were also analyzed by X-ray photoelectron spectroscopy (XPS)(ESCA3600Shimduz).2.4.Specific surface area and pore size distribution analysesAnalyses of physical characteristics of MOCS included spe-cific surface area,and pore size distributions.The specific sur-face area of MOCS and pore volumes were test using the nitrogen adsorption method with NOV A 1000High-Speed,Automated Surface Area and Pore Size Analyer (Quantachrome Corpora-tion,America)at 77K,and the BET adsorption model was used in the calculation.Calculation of pore size followed the method of BJH according to implemented software routines.2.5.Methods of adsorption studiesBatch adsorption studies were conducted by shaking the flasks at 120rpm for a period of time using a water bath cum mechanical shaker.Following a systematic work on the sorp-tion uptake capacity of Cu(II)and Pb(II)in batch systems were studied in the present work.The experimental process was as following:put a certain quantity of MOCS into conical flasks,then,added the solute of metals of copper or lead in single component system,vibrated sometime at a constant speed of 120rpm in a shaking water bath,when reached the sorption equilibrium after 180min,took out the conical flasks,filtrated to separate MOCS and the solution.No other solutions were provided for additional ionic strength expect for the effect of ionic strength.The concentration of the free metal ions in the filtrate was analyzed using flame atomic absorption spectrometer (AAS)(Aanalyst 300,Perkin Elmer).The uptake of the metal ions was calculated by the difference in their initial and final concentrations.Effect of pH (1.4–6.5),quantity of MOCS,contact time,temperature (288–318K)was studied.The pH of the solutions at the beginning and end of experiments was measured.Each experiment was repeated three times and the results given were the average values.2.5.1.Effect of contact time and temperature on Cu(II)and Pb(II)adsorptionA 2.0g l −1sample of MOCS was added to each 20ml of Cu(II)or Pb(II)solutions with initial concentration of Cu(II)0.315mmol l −1and Pb(II)0.579mmol l −1,respectively.The temperature was controlled with a water bath at the temperature ranged from 294to 318K for the studies.Adsorbent of MOCS and metal solution were separated at pre-determined time inter-R.Han et al./Journal of Hazardous Materials B137(2006)384–395387 vals,filtered and analyzed for residual Cu(II)and Pb(II)ionconcentrations.2.5.2.Effect of pH on the sorption of Cu(II)and Pb(II)byMOCSThe effect of pH on the adsorption capacity of MOCSwas investigated using solutions of0.157mmol l−1Cu(II)and0.393mmol l−1Pb(II)for a pH range of1.4–6.5at293K.A20g l−1of MOCS was added to20ml of Cu(II)and Pb(II)solu-tions.Experiments could not be performed at higher pH valuesdue to low solubility of metal ions.2.5.3.Effect of MOCS doseIt was tested by the addition of sodium nitrate and calciumnitrate to the solution of Cu(II)and Pb(II),respectively.The doseof adsorbents were varied from10to80g l−1keeping initial con-centration of copper0.157mmol l−1and lead0.393mmol l−1,respectively,and contact time was180min at the temperature of293K.2.5.4.Effect of ionic strength on Cu(II)and Pb(II)adsorptionThe concentration of NaNO3and Ca(NO3)2used rangedfrom0to0.2mol l−1.The dose of adsorbents were20g l−1,the initial concentration of copper0.157mmol l−1and lead0.393mmol l−1,respectively,and contact time was180min atthe temperature of293K.The data obtained in batch model studies was used to calculatethe equilibrium metal uptake capacity.It was calculated for eachsample of copper by using the following expression:q t=v(C0−C t)m(8)where q t is the amount of metal ions adsorbed on the MOCS at time t(mmol g−1),C0and C t the initial and liquid-phase concentrations of metal ions at time t(mmol l−1),v the volume of the aqueous phase(l)and m is the dry weight of the adsorbent(g).3.Results and discussion3.1.Mineralogy of manganese oxide coated sandThe samples of sand coated with manganese oxide were dark colored(brown–black)precipitates,indicating the presence of manganese in the form of insoluble oxides.The X-ray diffrac-tion spectrum(XRD)of the samples(data not shown)revealed that the manganese oxide were totally amorphous,as there was not any peak detected,indicative of a specific crystalline phase. SEM photographs in Fig.1were taken at10,000×magnifi-cations to observe the surface morphology of uncoated sand and manganese oxide coated sand,respectively.SEM images of acid-washed uncoated quartz sand in Fig.1(a)showed very ordered silica crystals at the surface.The virgin sand had a rela-tively uniform and smooth surface and small cracks,micropores or light roughness could be found on the sand -paring the images of virgin(Fig.1(a))and manganeseoxide Fig.1.SEM micrograph of sample:(a)sand;(b)manganese oxide coated sand. coated sand(Fig.1(b)),MOCS had a significantly rougher sur-face than plain sand and the coated sand surfaces were apparently occupied by newborn manganese oxides,which were formed during the coating process.Fig.1(b)also showed manganese oxides,formed in clusters,apparently on occupied surfaces.At the micron scale,the synthetic coating was composed of small particles on top of a more consolidated coating.In most regions individual particles of manganese oxide(diameter=2–3␮m) appeared to be growing in clumps in surface depressions and coating cracks.The amount of manganese on the surface of the MOCS,measured through acid digestion analysis,was approx-imately5.46mg Mn/g-sand.3.2.SEM/EDAX analysisThe elements indicated as being associated with manganese oxide coated were detected by the energy dispersive X-ray spec-trometer system(EDAX)using a standardless qualitative EDAX analytical technique.The peak heights in the EDAX spectra are proportional to the metallic elements concentration.The quali-tative EDAX spectra for MOCS(Fig.2(a))indicated that Mn,O,388R.Han et al./Journal of Hazardous Materials B137(2006)384–395Fig.2.EDAX spectrum of MOCS under:(a)adsorbed without copper and lead ion;(b)adsorbed copper ions;(c)adsorbed lead ions.Si,and K are the main constituents.These had been known as the principal elements of MOCS.EDAX analysis yielded indirect evidence for the mechanism of manganese oxide on the surface of MOCS.The peak of Si occurred in EDAX showed that man-ganese oxides do not covered a full surface of the MOCS.If the solid sample of MOCS caused a change of elemental con-stitution through adsorption reaction,it could be inferred that manganese oxide has already brought about chemical interac-tion with adsorbate.The EDAX spectrum for copper and lead system was illustrated in Fig.2(b and c).It could be seen that copper and lead ion became one element of solid sample in this spectrum.The reason was that copper and lead ions were chemisorbed on the surface of MOCS.Dot mapping can provide an indication of the qualitative abundance of mapping elements.The elemental distribution mapping of EDAX for the sample of MOCS and MOCS adsorbed copper or lead ions is illustrated in Fig.3.The bright points represented the single of the element from the solid sam-ple.A laryer of manganese oxide coating is clearly shown in the dot map for Mn in Fig.3(a),and a high density of white dots indicates manganese is the most abundant element.Results indicated that manganese oxide was spread over the surface of MOCS,and was a constituent part of the solid sample.The ele-ment distribution mapping of EDAX for the sample ofMOCS Fig.3.EDAX results of MOCS(white images in mapping represent the cor-responding element):(a)adsorbed without copper and lead ion;(b)adsorbed copper ions;(c)adsorbed lead ions.R.Han et al./Journal of Hazardous Materials B137(2006)384–395389Fig.4.XPS wide scan of the manganese oxide coated sand. reacting with copper and lead ions is illustrated in Fig.3(b and c).Copper or lead ions were spread over the surfaces of MOCS. Results indicated that manganese oxide produces chemical bond with copper or lead ions.Thus,copper or lead element was a constituent part of the solid sample.3.3.Surface characterization using the X-ray photoelectron spectroscopy(XPS)XPS analyses were performed on samples of MOCS alone and reacting with copper or lead ions.The wide scan of MOCS is presented in Fig.4.It can be noticed that the major elements constituent are manganese,oxygen,and silicon.Detailed spectra of the peaks are shown in Fig.5.Manganese oxides are generally expressed with the chemical formula of MnO x,due to the multiple valence states exhibited by Mn.Therefore,it is reasonable to measure the average oxidation state for a manganese mineral[37].The observation of the Mn 2p3/2peak at641.9eV and the separation between this and the Mn2p1/2peak of11.4eV indicates the manganese exhibited oxidation between Mn3+and Mn4+as shown from the auger plot,but it can be seen to show Mn4+predominantly from the Mn2p3/2peaks[38].The large peak in Fig.5(b)is a sum of the two peaks at 529.3and533.1eV,which can be assigned to O1s;a low bind-ing energy at529.7eV,which is generally accepted as lattice oxygen in the form of O2−(metal oxygen bond).This peak is characteristic of the oxygen in manganese oxides.The second peak at533.4eV can be assigned to surface adsorbed oxygen in the form of OH−[38].As seen the XPS spectra of the sample of MOCS reacting with copper,Fig.6(a)shows the binding energies of the observed photoelectron peaks of Cu2p3/2,2p1/2.The binding energy of the Cu2p3/2peak at a value of933.9eV shows the presence of copper(+2).The XPS spectra obtained after Pb(II)adsorption on MOCS is presented in Fig.6(b).Fig.6(b)shows that doublets charac-teristic of lead appear,respectively,at138.3eV(assigned to Pb 4f7/2)and at143.8eV(assigned to Pb4f5/2)after loadingMOCSFig.5.XPS detailed spectra of MOCS:(a)Mn2p3/2;(b)O1s.with Pb(II)solution.The peak observed at138.3eV agrees with the138.0eV value reported for PbO[39].This shows afixation of lead onto MOCS during the process.3.4.Specific surface area and pore size distribution analysesThe specific surface areas for sand and MOCS under un/adsorbed Pb(II)ions are summarized in Table1.Plain uncoated sand had a surface area of0.674m2g−1.A surface coating of manganese oxide increased the surface area of sand to0.712m2g−1,while average pore diameter decreased from 51.42to42.77˚A.This may be caused by the increase in both Table1Specific surface areas and average pore diameters for sand and various MOCSSurface area(m2g−1)Average pore diameter(˚A) Sand0.67451.42Unadsorbed a0.71242.77Adsorbed b0.55239.64Desorbed c0.70142.71a Without reacting with Pb(II)ions.b After reacting with Pb(II)ions.c After soaking with0.5mol l−1acid solution.390R.Han et al./Journal of Hazardous Materials B137(2006)384–395Fig.6.XPS detailed spectra of MOCS reacting with(a)copper;(b)lead. inner and surface porosity after adding the manganese oxides admixture.After reacting with Pb(II)ions,the pore size distribu-tion of MOCS had been changed,and parts of pores disappeared through the adsorption process.The results indicated the parts of pores were occupied with Pb(II)ions and average pore diameters decreased simultaneously,compared with unadsorbed MOCS, the surface area value of adsorbed MOCS is decreased,varying from of0.712to0.552m2g−1.Besides,pore size distribution of desorbed MOCS was similar to that of unadsorbed MOCS. The surface area of desorbed MOCS increased and average pore diameter also increased after regeneration with acid solution. The results indicated Pb(II)ions could be desorbed from the surface site of micropore and mesopores.3.5.Effect of contact time and temperature on Cu(II)andPb(II)adsorptionEffect of contact time and temperature on the adsorption of the copper(II)and lead(II)on MOCS was illustrated in Fig.7(a and b).The uptake equilibrium of Cu(II)and Pb(II) were achieved after180min and no remarkable changes were observed for higher reaction times(not shown in Fig.7).The shapes of the curves representing metal uptake versus time suggest that a two-step mechanism occurs.Thefirstportion Fig.7.Effect of contact time on Cu(II)and Pb(II)ions adsorption at pH4and various temperatures:(a)adsorption capacity vs.time;(b)adsorption percent vs.time(C0(Cu)=0.315mmol l−1,C0(Pb)=0.579mmol l−1).indicates that a rapid adsorption occurs during thefirst30min after which equilibrium is slowly achieved.Almost80%of total removal for both Cu(II)and Pb(II)occurred within60min.The equilibrium time required for maximum removal of Cu(II)and Pb(II)were90and120min at all the experimental temperatures, respectively.As a consequence,180min was chosen as the reac-tion time required to reaching pseudo-equilibrium in the present “equilibrium”adsorption experiments.Higher removal for cop-per and lead ions was also observed in the higher temperature range.This was due to the increasing tendency of adsorbate ions to adsorb from the interface to the solution with increasing temperature and it is suggested that the sorption of Cu(II)and Pb(II)by MOCS may involve not only physical but also chem-ical sorption.The metal uptake versus time curves at different temperatures are single,smooth and continuous leading to sat-uration,suggesting possible monolayer coverage of Cu(II)and Pb(II)on the surface of MOCS[40].3.6.Effect of pH on the sorption of Cu(II)and Pb(II)by MOCSIt is well known that the pH of the system is an important vari-able in the adsorption process.The charge of the adsorbate and the adsorbent often depends on the pH of the solution.The man-R.Han et al./Journal of Hazardous Materials B137(2006)384–395391 ganese oxide surface charge is also dependent on the solution pHdue to exchange of H+ions.The surface groups of manganeseoxide are amphoteric and can function as an acid or a base[41].The oxide surface can undergo protonation and deprotonationin response to changes in solution pH.As shown in Fig.8,the uptake of free ionic copper and leaddepends on pH,increasing with pH from1.4to5.1for Cu(II)and1.4to4.3for Pb(II).Above these pH levels,the adsorptioncurves increased very slightly or tended to level out.At low pH,Cu(II)and Pb(II)removal were inhibited possibly as result ofa competition between hydrogen and metal ions on the sorp-tion sites,with an apparent preponderance of hydrogen ions.Asthe pH increased,the negative charge density on MOCS sur-face increases due to deprotonation of the metal binding sitesand thus the adsorption of metal ions increased.The increase inadsorption with the decrease in H+ion concentration(high pH)indicates that ion exchange is one of major adsorption process.Above pH6.0,insoluble copper or lead hydroxide starts precip-itating from the solution,making true sorption studies impossi-ble.Therefore,at these pH values,both adsorption and precipita-tion are the effective mechanisms to remove the Cu(II)and Pb(II)in aqueous solution.At higher pH values,Cu(II)and Pb(II)inaqueous solution convert to different hydrolysis products.In order to understand the adsorption mechanism,the varia-tion of pH in a solution and the metal ions adsorbed on MOCSduring adsorption were measured,and the results are shown inFig.8.The pH of the solution at the end of experiments wasobserved to be decreased after adsorption by MOCS.Theseresults indicated that the mechanism by means of which Cu(II)and Pb(II)ion was adsorbed onto MOCS perhaps involved anexchange reaction of Cu2+or Pb2+with H+on the surface andsurface complex formation.According to the principle of ion-exchange,the more metalions that is adsorbed onto MOCS,the more hydrogen ions arereleased,thus the pH value was decreased.The complex reac-tions of Cu2+and Pb2+with manganese oxide may be writtenas follows(X=Cu,Pb and Y=Pb)[42]:MnOH+X2+ MnO−X2++H+(9)MnO−+X2+ MnO−X2+(10)Fig.8.Effect of pH on adsorption of Cu(II)and Pb(II)by MOCS.2(MnOH)+X2+ (MnO−)2X2++2H+(11)2(MnO−)+X2+ (MnO−)2X2+(12)MnOH+X2++H2O MnOXOH+2H+(13)MnOH+2Y2++H2O MnOY2OH2++2H+(14)Eqs.(9)–(14)showed the hydrogen ion concentration increasedwith an increasing amount of Cu(II)or Pb(II)ion adsorbed onthe MOCS surface.3.7.Effect of MOCS doseFig.9shows the adsorption of Cu(II)and Pb(II)as a functionof adsorbent dosage.It was observed that percent adsorptionof Cu(II)and Pb(II)increased from29to99%and19to99%with increasing adsorbent load from10to80g l−1,respectively.This was because of the availability of more and more bindingsites for complexation of Cu(II)ions.On the other hand,theplot of adsorption per unit of adsorbent versus adsorbent doserevealed that the unit adsorption capacity was high at low dosesand reduced at high dose.There are many factors,which can con-tribute to this adsorbent concentration effect.The most importantfactor is that adsorption sites remain unsaturated during theadsorption reaction.This is due to the fact that as the dosageof adsorbent is increased,there is less commensurate increasein adsorption resulting from the lower adsorptive capacity uti-lization of the adsorbent.It is readily understood that the numberof available adsorption sites increases by increasing the adsor-bent dose and it,therefore,results in the increase of the amountof adsorbed metal ions.The decrease in equilibrium uptake withincrease in the adsorbent dose is mainly because of unsaturationof adsorption sites through the adsorption process.The corre-sponding linear plots of the values of percentage removal(Γ)against dose(m s)were regressed to obtain expressions for thesevalues in terms of the m s parameters.This relationship is asfollows:for Cu(II):Γ=m s0.221+6.61×10−3m s(15)Fig.9.Effect of dosage of MOCS on Cu(II)and Pb(II)removal.。

Journal of Molecular Catalysis A:Chemical264(2007)153–161Preparation of mesoporous silica/polymer sulfonate composite materials Masahiro Fujiwara a,∗,Kumi Shiokawa a,Yingchun Zhu ba Kansai Center,National Institute of Advanced Industrial Science and Technology(AIST),1-8-31Midorigaoka,Ikeda,Osaka563-8577,Japanb Shanghai Institute of Ceramics,Chinese Academy of Sciences,Shanghai200050,People’s Republic of ChinaReceived21June2006;received in revised form22August2006;accepted9September2006Available online15September2006AbstractMesoporous silica/polymer sulfonate composite materials were prepared by simply mixing hexadecyltrimethylammonium bromide,polymer sulfonates and TEOS(tetraethoxysilane)in alkaline aqueous solution.Nafion and poly(sodium4-styrenesulfonate)were employed as polymer sulfonates.XRD patterns and nitrogen adsorption–desorption isotherms showed that the precipitates obtained had mesostructure similar to MCM-41.Especially,the crystallinity of hexagonal structure of composite materials synthesized with Nafion was high.From all the results obtained here, it is concluded that the polymer sulfonate resins might be incorporated in the wall framework of mesoporous silica matrix.However,when the excess amount of Nafion was mixed,the acid sites of Nafion were significantly lost in the obtained materials.These composite materials present new classes of organically modified mesoporous silicas,where organic polymers are incorporated in the framework of mesoporous silica.They were used as catalysts for␣-methylstyrene dimerization and Friedel–Crafts alkylation reaction of aromatics.©2006Elsevier B.V.All rights reserved.Keywords:Mesoporous silica;Nafion;Poly(styrenesulfonate);Nano-composite;Solid acid;␣-Methylstyrene dimerization1.IntroductionResearches on mesoporous silicas and related materials are importantfields of recent material science[1,2].Especially MCM-41and its analogues[2]are actively studied because of their high potentials for various applications.The function-alization of mesoporous silica with organic compounds began with the surface modification using silane compounds such as R-Si(OR )3[1,3].After this kind of approach,the framework modification using disilane compounds followed.These materi-als are often called periodic mesoporous organosilicas(PMOs) [4–6].For example,Inagaki and co-workers notified that ben-zene ring and analogues are completely incorporated into the framework of mesoporous silica materials,and that these mate-rials are effective acid catalysts after sulphonation[7].Another trend is the polymerizations in the pore voids of mesoporous materials:many researchers produced composite materials with the corresponding polymers by this method[8].Mesoporous composite materials,where an organic polymer is introduced into their“framework”,are also investigated.In2000,we briefly ∗Corresponding author.Tel.:+81727519253.E-mail address:m-fujiwara@aist.go.jp(M.Fujiwara).reported that Nafion resin,whose structure is illustrated inFig.1,was incorporated in the framework of M41S type ofmesoporous silica[9].This material was a unique catalyst for ␣-methylstyrene dimerization.Recently,another group devel-oped the composite materials with polyacrylate[10].In Fig.2,a classification of these composite materials of mesoporous sil-ica with organic components is proposed.Type(A)is surfacemodification using R-Si(OR )3compounds[1,3],and Type(B)isframework modification such as periodic mesoporous organosil-icas(PMOs)[4–7].Type(C)shows composite materials withpolymeric compounds in the pore voids[8].Composite meso-porous materials with polymers in the framework are namedType(D)here[9,10].In this paper,we wish to report further examination of thecomposite materials of mesoporous silica with Nafion resin.Thematrices of mesoporous materials are expected to offer orderednanostructures useful as solid support[11].Nafion resin is alsoa functional perfluorinated sulfonic acid polymer to be usedas acid catalyst[12]and as polymer electrolyte for fuel cellapplication[13].Composite materials of Nafion resin with amor-phous silica have been utilized in these technologies[14,15].The ordered nanostructures of Nafion and analogous resins withmesoporous silica matrix are expected to be useful for variousapplications.1381-1169/$–see front matter©2006Elsevier B.V.All rights reserved. doi:10.1016/j.molcata.2006.09.016154M.Fujiwara et al./Journal of Molecular Catalysis A:Chemical264(2007)153–161Fig.1.Structure of Nafion resin.2.Experimental2.1.Preparation of mesoporous silica/Nafion composite materialThe preparation procedure is considerably simplified from our previously reported method[9].Nafion solution commer-cial available was directly used and no hydrothermal treatment was performed.A typical synthesis of mesoporous silica/Nafion composite is following:5.0g of5%Nafion alcohol solution (from Aldrich)was added to200mL of the aqueous solution of NaOH(1.73g;43.25mmol)and hexadecyltrimethylammo-nium bromide(3.48g;9.55mmol),and this mixed solution was stirred for a few minutes.To this solution,16.69g(80mmol) of tetraethoxysilane(TEOS)was added dropwise for5min, and the resulting solution was further stirred for12h at room temperature.An as-synthesized sample thus obtained wasfil-tered,washed with sufficient amount of H2O and dried at80◦C for12h.Template was removed by refluxing with1M H2SO4 solution of EtOH(solid sample/EtOH solution=1g/150mL)for 12h.Thefiltered solid was refluxed again with pure EtOH(sam-ple/EtOH=1g/150mL)for12h,filtered,washed with H2O at room temperature and dried at80◦C for12h.2.2.Preparation of mesoporous silica/Nafion composite material from amorphous silica/Nafion compositeThe general preparation method of MCM-41type of meso-porous silica from porous amorphous silica is described in elsewhere[16].This procedure was applied to amorphous sil-ica/Nafion composite.The amorphous silica/Nafion composite used here was SAC-13purchased from Aldrich.To the aqueous solution of NaOH(0.15g;3.80mmol)with hexadecyltrimethy-lammonium bromide(0.37g;1.00mmol)in4mL of H2O,0.61g of SAC-13was added.After stirred for1h,the resulting mixture was placed in a stainless autoclave,sealed tightly and heated at 110◦C for24h under autogenous pressure.The following pro-cedures were similar to the above-mentioned process.2.3.Preparation of mesoporoussilica/poly(4-styrenesulfonate)composite materialTo the solution of NaOH(0.799g;19.98mmol)and hex-adecyltrimethylammonium bromide(1.582g;4.341mmol)in 90mL of H2O,0.412g of poly(sodium4-styrenesulfonate)dis-solved in10mL of H2O was added,and to this homogeneous solution8.403g(40.34mmol)of TEOS was mixed.The result-ing solution was stirred for2days at ambient temperature. The precipitate thus formed wasfiltered,washed with sufficient amount of H2O and air-dried.The following procedures were similar to the above-mentioned process.2.4.Preparation of mixture of Nafion with hexadecyltrimethylammonium bromideA5%alcohol solution of Nafion(7.656g;Nafion content: 0.353g;sulfonic acid equivalent:0.341mmol)was mixed with NaOH(0.038g:0.95mmol)in5mL of H2O.After removing solvent under reduced pressure,the residue was dissolved in a mixed solution of H2O(10mL)and EtOH(10mL).Finally hexadecyltrimethylammonium bromide(0.125g;0.343mmol) was added.After about1day,white precipitate obtained was filtered and air-dried.2.5.Preparation of mixture of poly(4-styrenesulfonate)with hexadecyltrimethylammonium bromideThe solution of0.412g of poly(sodium4-styrenesulfonate) in6mL of H2O was mixed with the aqueous solution(50mL)of hexadecyltrimethylammonium bromide(0.728g;2.00mmol). The white precipitate was formed in a few minutes.After stirring for1h,the white precipitate wasfiltered and dried at60◦C. 2.6.Product characterizationsXRD patterns were recorded with a MAC Science MXP3V diffraction apparatus with Nifiltered Cu K␣radiation (λ=0.15406nm).N2adsorption-desorption isotherms were obtained at−196◦C(in liquid N2)using a Bellsorp Mini instru-ment(BEL JAPAN Inc.).BJH calculation was performed toesti-Fig.2.Conceptual schemes of composite materials of mesoporous silica with organic components.M.Fujiwara et al./Journal of Molecular Catalysis A:Chemical264(2007)153–161155mate the mesopore size using adsorption branches of isotherms. Elemental analyses of silicon were performed by the alkali fusion-gravimetric method according to JIS G1212(Japanese industrial standard).Elemental analyses offluorine were car-ried out with the lanthanum-alizarin complexone method using a Shimadzu UV-1600photometry apparatus after the extraction of alkali fusion method.Elemental analyses of carbon were per-formed by the common combustion gas quantification method. Thermogravimetric analyses(TGA)were performed on a Shi-madzu TGA-50apparatus.All samples were held in a platinum sample holder and were heated under air from room temperature to800◦C at the rate of5◦C/min.FT-IR spectra were mea-sured on a Perkin-Elmer Spectrum One spectrometer.Transmit-tance electron microscope(TEM)images were obtained using a JEM-2100F(JEOL)high-resolution transmissionfield emis-sion electron microscope(HRTEM)operated at300kV.The acid capacities of composite materials were estimated by the titra-tion method.The composite materials were immersed in0.1M of aqueous solution of NaCl,and the acid amounts of the ion-exchanged solutions thus obtained were analyzed by titrating with0.01M NaOH using phenolphthalein as indicator.2.7.Catalytic reactionsThe experimental procedure of␣-methylstyrene(AMS) dimerization was described in our previous paper[9].A com-petitive Friedel–Crafts reaction of toluene and p-xylene with benzyl alcohol was performed by the mixed solution of toluene (0.92g,10mmol),p-xylene(1.05g,10mmol)and benzyl alco-hol(0.22g,2mmol)in the presence of a catalyst(0.05g)at 90◦C for7h with vigorous stirring.Afterfiltering the catalyst, thefiltrate was analyzed by a capillary GC.3.Results and discussion3.1.Synthesis of mesoporous silica/Nafion composite materialComposite materials made of mesoporous silica and Nafion resin were prepared by a procedure modified formesoporous Fig.3.XRD patterns of as-synthesized and template-free(solvent extracted) mesoporous silica/Nafion composite materials.(A)As-synthesized MCM/ Nafion-1.(B)As-synthesized MCM/Nafion-2.(C)Template-free MCM/Nafion-1.(D)Template-free MCM/Nafion-2.silica synthesis.TEOS was added to a homogeneous alkaline solution of hexadecyltrimethylammonium bromide with Nafion resin.After stirring at room temperature,as-synthesized com-posite materials of mesoporous silica and Nafion resin were obtained as a white precipitate.Although hydrothermal treat-ment in an autoclave was given in our previous paper[9],we found that this hydrothermal treatment is not essential for the synthesis after the publication of the paper.The surfactant as template was removed by refluxing in H2SO4–EtOH solution (1M of H2SO4).H2SO4is expected to contribute to both the regeneration of sulfonic acid sites in Nafion polymer resin and the surfactant removal.The sample names and the profiles of composite materials are summarized in Table1.The data of a composite material prepared under hydrothermal conditions[9] are also included in Table1as MCM/Nafion-H.Fig.3shows XRD patterns of two composite materials(MCM/Nafion-1andTable1Properties of various mesoporous silica/polymer composites materialsSample Starting ratio d100a SSA b(m2/g)PV c(cm3/g)PPD d(nm) g/mol e wt%f2θnmMCM/Nafion-1 3.13 5.21 2.32(2.34) 3.81(3.77)1239 1.12 2.52 MCM/Nafion-2 6.7211.18 2.20(2.22) 4.01(3.98)1211 1.26 2.75 MCM/Nafion-314.8724.76 2.20(2.28) 4.01(3.87)3330.26 2.52 MCM/Nafion-H g 6.5810.96 2.25(2.35) 3.92(3.76)918 1.00 2.75 MCM/PSS8.8815.19 2.14(2.24) 4.13(3.94)8380.66 2.75a d100:X-ray diffraction(100)interplanar spacing.In parentheses,as-synthesized sample.b BET specific surface area.c Primary mesopore volume calculated from adsorption branch of BJH pore size distribution curve.d Peak pore diameter from adsorption branch of BJH pore size distribution curve.e Starting ratio of polymer(as acid type)and TEOS;gram of polymer/molar of TEOS.f Estimated weight percentage when polymer is completely incorporated in solid material and all TEOS converts into silica(SiO2).g MCM/Nafion composite material we reported[9].156M.Fujiwara et al./Journal of Molecular Catalysis A:Chemical264(2007)153–161Fig.4.Infrared spectra of as-synthesized(A)and template-free(B)MCM/ Nafion-1composite material.MCM/Nafion-2)in as-synthesized and template-free forms. XRD patterns indicated the formation of the hexagonal structure characteristically observed in the MCM-41type of mesoporous silica[2].The peaks assigned to d110,d200and d210interpla-nar spacings were found besides those from d100interplanar ones in all four samples.The surfactants were removed success-fully by the extraction using H2SO4in EtOH with the hexagonal structure maintained.Template removal by calcination was not performed to avoid thermal decomposition of Nafion resin.The peaks derived from the hexagonally ordered structure became stronger after the template removal in both cases(MCM/Nafion-1and MCM/Nafion-2).Infrared spectra of as-synthesized and template-free (extracted)samples are shown in Fig.4.In the as-synthesized sample,strong absorptions of the surfactant were observed approximately at2929and2850cm−1(Fig.4A).These absorp-tions disappeared after the treatment with H2SO4(Fig.4B), indicating the complete removal of the surfactant.On the other hand,the absorptions of C–F stretching modes of Nafion resin at1210and1160cm−1were not found in either spectra,while they are detected in amorphous silica/Nafion composite[14a]. It seems that in the case of amorphous silica/Nafion compos-ite,the contact time of Nafion resin with alkaline solution is comparatively short(the preparation solution gels immediately), preventing the serious decomposition of C–F bonds[14a].In our case,Nafion resin was dissolved in the high alkaline solution for a long time,resulting in critical degradation.The TEM images of MCM/Nafion-1are shown in Fig.5.The ordered structure(hexagonally arranged)was confirmed from the layered lines in the solid.The distance between layers is estimated to be3.0–3.8nm,approximately according with that from XRD patterns.The nitrogen adsorption–desorption isotherms of the template-free samples are shown in Fig.6.Both samples, MCM/Nafion-1and MCM/Nafion-2,indicated the typical type IV isotherms(IUPAC)of ordered mesoporous silica materi-als.The pore size of MCM/Nafion-2was larger than that of MCM/Nafion-1.These results were consistent with the d100 interplanar spacings from XRD patterns(Fig.3).Specific sur-face areas(BET surface area)and pore volumes of both samples were over1000m2/g and1cm3/g,respectively.These data were at the level similar to the MCM-41type of mesoporous sil-icas[1,2],and considerably higher than those of amorphous silica/Nafion composite materials[15].These properties are similar to those of the sample prepared under hydrothermal con-ditions(MCM/Nafion-H)in our previous paper[9].Therefore, a simpler preparation method using the direct use of commer-cial reagent under ambient conditions proved to be applicable. However,when more than20wt%of Nafion resin was added to the starting solution,the ordered structure of the corresponding composite material was considerably collapsed(MCM/Nafion-3).The peak at2.28in2θobserved in the as-synthesized sample (Fig.7A)indicated its moderately ordered structure.However, this peak almost disappeared after the removal of template (Fig.7A),showing the destruction of the ordered structure. In Fig.7B,the nitrogen adsorption–desorption isotherm and the pore size distribution estimated from the BJH method of template-free MCM/Nafion-3are presented.The porosity of this sample was poor and the peak of the pore diameterwas Fig.5.TEM images of MCM/Nafion-1(template-free).M.Fujiwara et al./Journal of Molecular Catalysis A:Chemical264(2007)153–161157Fig.6.(A)Nitrogen adsorption–desorption isotherms of mesoporous silica/Nafion composite materials.( )Adsorption branch of MCM/Nafion-1;( )desorption branch of MCM/Nafion-1;( )adsorption branch of MCM/Nafion-2;(᭹)desorption branch of MCM/Nafion-2.(B)Pore size distributions estimated from the adsorption branches of the isotherms by BJH method.( )MCM/Nafion-1;( )MCM/Nafion-2.broad.The specific surface area and the pore volume decreased to333m2/g and0.259cm3/g,respectively.Thus,the addition of excess amount of Nafion resin inhibited the formation of ordered structure.Another approach to the preparation of the composite mate-rial was attempted by using an amorphous silica/Nafion com-posite.It is well known that porous amorphous silica can be transformed into mesoporous MCM-41type material in the pres-ence of surfactant in alkaline solution[2,16].An amorphous silica/Nafion composite material commercially available(SAC-13;Nafion content:approximately13wt%)was immersed in an alkaline solution dissolving hexadecyltrimethylammonium bromide.This solution system was placed in an autoclave to be hydrothermally treated by reacted at115◦C for24h[16]. The XRD patterns of as-synthesized and template-free samples thus obtained are shown in Fig.8.The crystallinity of the as-synthesized sample was poor,and after the removal of template the hexagonal structure nearly collapsed.Even in this case,a comparatively high content of Nafion resin(13wt%)is thought to prevent the formation of ordered structure in the case of MCM/Nafion-3.3.2.Analyses of composition of mesoporous silica/Nafion composite materialThe contents of Nafion resin in these composite materials were analyzed by various methods.The results of TGA measure-ment of these composite materials are listed in Table2.Nafion resin is thermally decomposed from150to600◦C[17],and the weight decrease of pure mesoporous silica(without Nafion resin)we prepared was measured4.78%due to the thermal dehydration of silanols in this temperature range.The corrected values of the weight decreases by this blank measurement are shown in the parentheses.Although there are no direct propor-tional relationships between the starting contents of Nafion and the weight decreases,combustible Nafion contents increased from MCM/Nafion-1to MCM/Nafion-3.A similar tendency was observed in the elemental analysis shown in Table2.The Fig.7.(A)XRD patterns of as-synthesized and template-free(solvent extracted)MCM/Nafion-3.(B)Nitrogen adsorption–desorption isotherm and the pore size distribution by BJH method from adsorption branch(in inset)of MCM/Nafion-3.158M.Fujiwara et al./Journal of Molecular Catalysis A:Chemical264(2007)153–161Fig.8.XRD patterns of as-synthesized(A)and template-free(B)mesoporous silica/Nafion composite material obtained from an amorphous silica/Nafion composite(SAC-13).carbon andfluorine contents increased with the starting Nafion resin contents.It should be noted that thefluorine contents were low in these composite materials,although the weight ratios of fluorine to carbon must be approximately3.2according to the chemical formula of Nafion(Fig.2)[14a].These lower con-tents offluorine indicated that carbon–fluorine bonds in Nafion resin were significantly cleaved during the preparation process, because aliphatic perfluoro group is known to be unstable under basic conditions(although aromatic C–F bond is reported to be tolerant in alkaline solution)[11a].No observation of C–F bonds in infrared spectra(Fig.4),which can be observed in amorphous silica/Nafion composite[14a],was likely to result from the decrease in thefluorine content in the resin.The acid capacities of these composite materials were estimated by the cation exchange method with NaCl[14].The acid equivalents are also summarized in Table2.The acid content of pure Si-MCM-41(with Nafion resin)was under 0.001mequiv.H+/g.Pure Nafion resin(NR-50)and its compos-ite material with amorphous silica(SAC-13;Nafion content: 13wt%)is reported to have0.89or0.14mequiv.H+/g of acid capacities,respectively[14].In the parentheses of Table2, the weight percentages of Nafion in the composite materials calculated from the measured acid capacities are listed on the assumption that all sulfonic acid sites of Nafion resin are active. The calculated Nafion content of MCM/Nafion-1(5.92wt%) from acid capacity was reasonably consistent with the estimated values from both starting ratio and TGA measurement.On the other hand,in the case of MCM/Nafion-2,the Nafion content estimated from the acid capacity(17.02wt%)was in discord with other results.Furthermore,the acid capacity of MCM/Nafion-3was approximately0.003mequiv.H+/g.From these results,it was concluded that high contents of Nafion in the composite materials led to the some decomposition of Nafion resin,while no significant changes were observed in the case of low Nafion contents.3.3.Synthesis of mesoporous silica/poly-sulfonate composite materialThe formation of composite materials of mesoporous silica with polyacrylate was recently claimed in a report[10],where a procedure analogous to ours was used.We also studied the prepa-ration of a composite material made of mesoporous silica and another poly-sulfonate.Poly(4-styrenesulfonic acid)sodium salt was employed for the synthesis.In a similar manner to Nafion, TEOS was added to the mixed alkaline solution of poly(sodium 4-styrenesulfonate)(PSS)and hexadecyltrimethylammonium bromide,forming a composite material(MCM/PSS)after stir-ring.The XRD patterns of the as-synthesized and template-free samples are shown in Fig.9A.Although the peak intensities of these two patterns were lower than those of composite materials with Nafion,an ordered structure in the nano-level was observed. In the as-synthesized MCM/PSS,peaks assigned to d110,d200 and d210interplanar spacings were found as well as d100inter-planar one.Those peaks are not so clear in the template-free MCM/PSS,and its pore structure might be a wormhole like one[18].The nitrogen adsorption–desorption isotherms of this MCM/PSS composite material presented in Fig.9B are basi-cally type IV.The peak pore diameter was found at2.75nm (in inset).Thus,a poly(styrenesulfonate)polymer can be suc-cessfully introduced into mesoporous silica material as well as Nafion resin.Table2Results of elemental analysis,TGA and acid capacity of mesoporous silica/polymer composite materialsSample Stating ratio(wt%)a Elemental analysis(wt%)b TGA(%)c Acid capacity(mequiv.H+/g)dC Si FMCM/Nafion-1 5.21 2.7539.1 1.659.06(4.27)0.0527(5.92)MCM/Nafion-211.18 2.8939.3 1.7110.75(5.97)0.1515(17.02)MCM/Nafion-324.76 3.9236.8 5.8713.99(9.21)0.003(0.3)a Starting weight composition of Nafion calculated from carbon in Nafion and silicon in TEOS,regarding as Nafion formula are n=7and m=1in Fig.1.b Elemental analyses of C,Si and F were performed by common combustion gas quantification method,alkali fusion-gravimetric method or lanthanum-alizarin complexone method,respectively.c Percentage of weight loss from150to600◦C.In parentheses,the corrected value by deducting the weight decrease(4.78%)by the dehydration of mesoporous silica prepared without Nafion is noted.d Acid capacity estimated from the titration of ion-exchanged solution from NaCl using NaOH solution.In parentheses,the weight percent of Nafion calculated from this acid capacity using the pure Nafion resin acid capacity[14a],0.89mequiv.H+/g(on the supposition that all sulfonic acid sites are active).M.Fujiwara et al./Journal of Molecular Catalysis A:Chemical 264(2007)153–161159Fig.9.(A)XRD patterns of as-synthesized and template-free (solvent extracted)mesoporous silica/poly(sodium 4-styrenesulfonate)composite material (MCM/PSS).(B)Nitrogen adsorption–desorption isotherm and the pore size distribution by BJH method from adsorption branch (in inset)of MCM/PSS.3.4.Mechanistic discussion on the formation ofmesoporous silica/poly-sulfonate composite materials It is well known that the polymer electrolyte such as ion-exchange resin and surfactant readily form their complex by their ionic interaction [19].When sodium polyacrylate or poly(4-styrenesulfonate)was mixed with hexadecyltrimethy-lammonium bromide in aqueous solution,their complexes were instantly produced as precipitates.On the other hand,sodium salt of Nafion resin obtained by neutralization with sodium hydroxide scarcely afforded the precipitate with the surfactant in aqueous solution.Only after the considerable evaporation of solvent,a white viscous solid was obtained.Fig.10shows XRD patterns of the complexes obtained from sodium salt of Nafion or sodium poly(4-styrenesulfonate)withhexadecyltrimethylam-Fig.10.XRD patterns of precipitated solids from Nafion resin (A)or sodium poly(4-styrenesulfonate)(B)with hexadecyltrimethylammonium bromide.monium bromide.In the XRD pattern of the complex from sodium poly(4-styrenesulfonate)and the surfactant,a sharp peak at 2.11in 2θ(interplanar space:4.18nm)was found (B in Fig.10).These kinds of XRD pattern often observed in the com-plexes of polyelectrolytes and surfactants indicate the formation of lamellar structure complexes of polymer electrolytes with sur-factants [19d].On the other hand,in the case of the complex from Nafion resin and surfactant,no clear peak was observed in the XRD pattern (A in Fig.10),indicating that Nafion forms no ordered complex with cationic surfactant.A broad peak found at 2.22in 2θ(interplanar space:3.98nm)is likely to be derived from the cluster structure of sulfonic acid parts of Nafion resin [20].This cluster structure might restrict the formation of the complex with surfactant.In Fig.11,a possible formation mechanism of composite material consisting of mesoporous silica and polymer sulfonate is displayed.Two routes of composite materials formation are assumed.In the case of MCM/PSS synthesis,some layered phases of poly(styrenesulfonate)and surfactant are formed at first.The hydrolysis of TEOS to silica occurs in this solu-tion.With the progress of the condensation of silanols (Si–OH)to siloxane bonds (Si–O–Si),the hexagonal structure by the influences of surfactant is formed gradually.However,the lay-ered structure of poly(styrenesulfonate)and surfactant is com-paratively strong so as to restrict the transformation of the layered structure to the hexagonal one (route A).The lower crystallinity of MCM/PSS is thought to be caused from this effect.On the other hand,complex compounds are scarcely formed from Nafion resin and surfactant,not suppressing the above-mentioned transformation and the fabrication of hexag-onal structure (route B).It is thought that the electrostatic interaction between the sulfonate group of Nafion and cationic surfactant compels the mixing of Nafion resin in aqueous phase as shown in route B of Fig.11,when the amount of Nafion is not overabound.It is not sure that Nafion resin bearing highly hydrophobic perfluoro main chain is incorporated into the aque-ous phase of the mixed solution.However,the low fluorine contents in MCM/Nafion composite materials confirmed by the160M.Fujiwara et al./Journal of Molecular Catalysis A:Chemical 264(2007)153–161Fig.11.Expected mechanisms of mesoporous silica/polymer sulfonate composite materials.elemental analysis suggested that the reaction of carbon–fluorine bond proceeds to eliminate fluorine in high alkaline solution.The main chains of Nafion resin become more hydrophilic by this reaction,increasing the affinity for silica matrix.Finally,the well-defined hexagonal structure of MCM/Nafion compos-ite materials is obtained in the case of low loading of Nafion resin.3.5.Catalytic Friedel–Crafts reaction by mesoporous silica/Nafion composite materialsWe have previously shown the unique behavior of meso-porous silica/Nafion composite materials for ␣-methylstyrene (AMS)dimerization.Representative results are listed in ing this catalyst,intermediate products (products 1and 2)are predominantly obtained and the further reaction (intramolecular Friedel–Crafts reaction)to form an indan deriva-tive (product 3)is inhibited,while the product 3was yielded effectively by amorphous silica/Nafion composite (SAC-13)[9].A competitive Friedel–Crafts type reaction of toluene and p -xylene with benzyl alcohol was examined using MCM/Nafion-1and SAC-13(Fig.12B).While no selectivity for the benzy-lation of toluene or p -xylene was observed in the reaction by SAC-13,the reaction of p -xylene occurred preferably in the case of MCM/Nafion-1catalyst.These results indicated that Friedel–Crafts reaction catalyzed by MCM/Nafion-1is more influenced by the substituents on the benzene ring than that by SAC-13.p -Xylene with two electron-donating groupsisFig.12.Results of ␣-methylstyrene (AMS)dimerization (A)and competitive Friedel–Crafts type reaction of toluene and p -xylene with benzyl alcohol (B).。

Infrared spectroscopic and X-ray diffraction characterizationof the nature of adsorbed arsenate on ferrihydriteYongfeng Jiaa,*,Liying Xu a ,Xin Wang a ,George P.Demopoulosb,*aInstitute of Applied Ecology,Key Laboratory of Terrestrial Ecological Process,Chinese Academy of Sciences,Shenyang 110016,ChinabDepartment of Mining,Metals and Materials Engineering,McGill University,Montreal,QC,Canada H3A 2B2Received 30March 2006;accepted in revised form 14December 2006;available online 30January 2007AbstractFourier transformed infrared (FTIR)spectroscopy was used to characterize arsenate–ferrihydrite sorption solids synthe-sized at pH 3–8.The speciation of sorbed arsenate was determined based on the As–O stretching vibration bands located at 650–950cm À1and O–H stretching vibration bands at 3000–3500cm À1.The positions of the As–O and O–H stretching vibra-tion bands changed with pH indicating that the nature of surface arsenate species on ferrihydrite was strongly pH dependent.Sorption density and synthesis media (sulfate vs.nitrate)had no appreciable effect.At acidic pH (3,4),ferric arsenate surface precipitate formed on ferrihydrite and constituted the predominant surface arsenate species.X-ray diffraction (XRD)analyses of he sorption solids synthesized at elevated temperature (75°C),pH 3clearly showed the development of crystalline ferric arsenate (i.e.scorodite).In neutral and alkaline media (pH 7,8),arsenate sorbed as a bidentate surface complex (in both pro-tonated B FeO 2As ðO ÞðOH ÞÀand unprotonated B FeO 2As ðO Þ22Àforms).For the sorption systems in slightly acidic media (pH 5,6),both ferric arsenate and surface complex were probably present on ferrihydrite.It was further determined that the incor-porated sulfate in ferrihydrite during synthesis was substituted by arsenate and was more easily exchangeable with increasing pH.Ó2007Elsevier Ltd.All rights reserved.1.INTRODUCTIONFerrihydrite is a poorly ordered hydrous iron oxide commonly present in low-temperature geochemical pro-cesses.It is widely occurring in surface environments,e.g.in soils,lake and river sediments and water columns (Way-chunas et al.,1993;Cornell and Schwertmann,1996;Jam-bor and Dutrizac,1998).Arsenate is an important form of arsenic in both natural water systems and mineral pro-cessing tailings.Ferrihydrite shows strong affinity for arse-nate at mineral–water interfaces.Adsorption on ferrihydrite is an important factor controlling transport,fate and bioavailability of arsenic in soils,groundwaterand surface water systems (Smedley and Kinniburgh,2002).Ferrihydrite is also a common product in many hydrometallurgical operations such as the coprecipitation process of iron with arsenate as well as in acid mine drain-age (Jambor and Dutrizac,1998;Carlson et al.,2002).Hence the adsorption of arsenate on ferrihydrite plays an important role in the removal and immobilization of ar-senic from industrial effluents as well as the fate of arsenate in tailings impoundment.Adsorption of arsenate on ferrihydrite involves ligand exchange with H 2O and/or OH Àon the substrate surface (Jain et al.,1999).Factors influencing adsorption of arse-nate on ferrihydrite include medium pH,type and con-centration of co-ions,initial Fe/As molar ratios etc.pH is one of the most important factors that control aqueous arsenate speciation and surface functional groups of hy-drous oxides,consequently influencing the macroscopic characteristics of adsorption process and the microscopic characteristics such as bonding modes and the nature of0016-7037/$-see front matter Ó2007Elsevier Ltd.All rights reserved.doi:10.1016/j.gca.2006.12.021*Corresponding authors.Fax:+862483970436(Y.Jia),+15143984492(G.P.Demopoulos).E-mail addresses:yongfeng.jia@ (Y.Jia),george.demopoulos@mcgill.ca (G.P.Demopoulos)./locate/gcaGeochimica et Cosmochimica Acta 71(2007)1643–1654adsorbed arsenate on the surface of ferrihydrite(Mas-scheleyn et al.,1991;Hsia et al.,1992;Fuller et al., 1993;Bowell,1994;Wilkie and Hering,1996;Raven et al.,1998;Jain et al.,1999;Meng et al.,2000;Grafe et al.,2002;Dixit and Hering,2003).The degree of pro-tonation of arsenate anion in aqueous solution is a func-tion of pH with p K a1=2.3,p K a2=6.8and p K a3=11.6 (Goldberg and Johnston,2001),resulting in arsenate spe-cies varying from H3AsO4,H2AsO4À,HAsO42À,toAsO43Àwhen pH increases from acidic region to alkalineregion(Myneni et al.,1998;Raven et al.,1998;Goldberg and Johnston,2001).On the other hand,the presence and the density of surface groups of ferrihydrite,i.e. H2O,OHÀ,are also strongly pH dependent.The point of zero charge(PZC)is approximately8.5(Jain et al., 1999;Goldberg and Johnston,2001).Hence,the modes of complexation of arsenate anions with ferrihydrite by replacing surface hydroxyl groups and/or waters are lar-gely controlled by the pH of reaction medium.Ligand exchange in the mode of bidentate binuclear in-ner-sphere complexation is the widely accepted mechanism of the adsorption of arsenate on iron oxides.It has been proposed based on infrared(Harrison and Berkheiser, 1982)and extended X-ray absorptionfine structure(EX-AFS)(Waychunas et al.,1993)analyses and confirmed to be the dominant interaction mode of arsenate–ferrihydrite and arsenate–goethite systems(Lumsdon et al.,1984;Man-ceau,1995;Sun and Doner,1996;Waychunas et al.,1996; Fendorf et al.,1997;Foster et al.,1998;Myneni et al.,1998; O’Reilly et al.,2001;Roddick-Lanzilotta et al.,2002;Sher-man and Randall,2003;Arai et al.,2004;Cance`s et al., 2005;Waychunas et al.,2005).However,most of the stud-ies were conducted at neutral to alkaline pH.The applica-bility of the conclusions to the acidic arsenate–ferrihydrite adsorption system may be questionable.Moreover,most of the studies have dealt mainly with characterizing the bonding mechanism between arsenate anions and surface iron polyhedra without identifying arsenate species on the surface of iron oxide.Based on macroscopic measurements of the adsorption process,arsenate was adsorbedas B FeO2AsðOHÞ2;B FeO2AsðOÞðOHÞÀ;B FeO2AsðOÞ22Àon the surface of ferrihydrite at mildly acidic and alkaline pH(Jain et al.,1999).In a recent work on direct character-ization of arsenate coordination on mineral(portlandite, gibbsite,ettringite,Fe-oxyhydroxides)surfaces using FTIR, both protonated and unprotonated arsenate species(i.e.HAsO42Àand AsO43Àwere present on the goethite surfaceat alkaline pH(Myneni et al.,1998).When arsenate was ad-sorbed on schwertmannite and ferrihydrite at acidic pH(i.e. pH3),surface precipitates were proposed to form and were termed as ferric hydroxyarsenate(FeOHAs)(Carlson et al., 2002).Evidence for surface precipitation of phosphate on goethite has been observed(Ler and Stanforth,2003).Sim-ilarly,surface precipitation of ferric arsenate on ferrihydrite is likely to occur in addition to bidentate binuclear com-plexation according to the XRD and Raman spectroscopic evidence we reported recently(Jia et al.,2006).The objective of this paper was to provide further evi-dence of surface precipitation of ferric arsenate on ferrihy-drite.This was done via characterization of the interactions between arsenate and ferrihydrite in terms of bonding modes and surface arsenate species as a function of pH and coverage density by Fourier transformed infrared spec-troscopy(FTIR)and evolution of crystallinity at elevated temperature(75°C)by X-ray diffraction(XRD)analysis. The effect of ferrihydrite synthesis media(NO3Àvs.SO42À)was also evaluated since it significantly influenced arsenate adsorption capacity(Jia and Demopoulos,2005). Poorly crystalline ferric arsenate was used as reference material in the study.Moreover,relatively high arsenic con-centration solutions were used in this study since this is the case in important hydrometallurgical operations where arsenic removal is practiced.2.MATERIALS AND METHODS2.1.Synthesis of poorly crystalline ferric arsenatePoorly crystalline ferric arsenate was synthesized at 21°C by adjusting a0.02M As(V)/0.02M Fe(III)solution (as sodium arsenate and ferric sulfate)from initial pH1.3to 1.8with NaOH solution and maintained at that pH for1h (Jia et al.,2006).The resultant solid was separated byfiltra-tion,washed with de-ionized water(pH2)and vacuum-dried at60°C.The chemical formula(Fe1.02AsO4Æ2.4H2O) was determined by digestion with hydrochloric acid fol-lowed by ICP-AES analysis.2.2.Synthesis of arsenate–ferrihydrite sorption samplesTwo-line ferrihydrite samples were synthesized at21°C using a slightly modified procedure from that reported in the literature(Schwertmann and Cornell,1991).Both ni-trate(Fe(NO3)3Æ9H2O)and sulfate(Fe2(SO4)3Æ5H2O)salts were used as the sources of ferric iron[Fe(III)].Briefly, the Fe(III)solution was prepared by dissolving ferric ni-trate or ferric sulfate in de-ionized water.The pH of the solution was raised to$7.5in about5min using1M NaOH solution and maintained at that pH for1h with the slurry mechanically agitated vigorously.The ferrihy-drite samples synthesized from sulfate and nitrate media were termed as‘‘sulfate–ferrihydrite’’and‘‘nitrate–ferrihy-drite’’,respectively,for simplicity.The prepared ferrihydrite slurry was adjusted to differ-ent pH between3and8with NaOH and HNO3and al-lowed to equilibrate for1h.Arsenate solution was introduced into the ferrihydrite slurry from a burette over a10-min period with the slurry mechanically stirred moder-ately.The pH was controlled by addition of NaOH and/or HNO3solution and allowed to equilibrate at21°C for 2weeks.The volume of adsorption slurry was500mL for all experiments and the concentration of Fe(III)in the slur-ry system was4g/L.At each pH,three initial Fe/As molar ratios(i.e.Fe/As=2,4and8)were applied.Arsenate–fer-rihydrite sorption samples were also synthesized at75°C, Fe/As molar ratio of2and4.The pH of the slurry was con-trolled constant at pH3throughout the sorption reaction. Samples were taken at1day,3day,1week,2weeks and 2months of reaction time.The synthesized arsenate–fer-rihydrite sorption products werefiltered,DI water-rinsed1644Y.Jia et al./Geochimica et Cosmochimica Acta71(2007)1643–1654and vacuum-dried at60°C.The equilibrium concentration of arsenic was determined by ICP-AES analysis.2.3.FTIR analysisThe infrared spectra of the samples were obtained on a Bio-Rad FTS60Fourier Transformed Infrared Spectrome-ter with a MCT liquid nitrogen cooled detector.The KBr/ sample discs were prepared by mixing0.5%offinely ground samples in KBr.The sample chamber was purged by N2gas for10min before scans were started.The measurement res-olution was set at4cmÀ1and the spectra were collected in the range of400–4000cmÀ1with200co-added scans.2.4.X-ray diffraction(XRD)analysisThe powder XRD patterns were collected on a Rigaku D/Max2500PC X-ray diffractometer with graphite mono-chromated CuK a1radiation.The powder samples were scanned from10to90°2h with increments of0.02°2h.3.RESULTS AND DISCUSSION3.1.Effect of pH on the nature of adsorbed arsenateAqueous arsenate species have no direct bearing on the surface arsenate species adsorbed on mineral surfaces (Myneni et al.,1998).However,the effect of complexation of arsenate ions on oxide surfaces is similar to that of pro-tonation of aqueous arsenate species.A brief discussion on infrared absorption of aqueous arsenate can assist with understanding the infrared characteristics of surface species (Myneni et al.,1998;Goldberg and Johnston,2001;Rod-dick-Lanzilotta et al.,2002).The free arsenate anion,AsO43À,is present in highly alkaline(p K a3=11.6)aqueoussolution and belongs to T d symmetry.Only m3and m4funda-mental bands are infrared active in this form.The infrared spectrum of an AsO43Àdominated solution exhibits a major band at792cmÀ1(Roddick-Lanzilotta et al.,2002).Uponprotonation or complexation with metal cations,the sym-metry decreases and splitting of the m3band occurs(Harri-son and Berkheiser,1982;Myneni et al.,1998).AqueousHAsO42Àspecies belong to the C3v symmetry.Its infrared spectrum shows two broad bands at859and689cmÀ1, the latter was assigned to stretching vibration of As–OH (Myneni et al.,1998).Curvefitting of the former band gave two bands at865and846cmÀ1,which were assigned to asymmetric and symmetric stretching vibration of uncom-plexed As–O,respectively(Myneni et al.,1998).Poorly crystalline ferric arsenate was used as reference material in this study to identify the possible occurrence of a surface precipitate of arsenate on ferrihydrite.This compound is not well defined and often termed loosely as amorphous ferric arsenate(Krause and Ettel,1989)or amorphous scorodite(Langmuir et al.,1999),because it possesses similar bonding structures to crystalline ferric arsenate,i.e.scorodite(FeAsO4Æ2H2O).It is an unstable arsenate phase with increasing pH and tends to convert to ferrihydrite(Krause and Ettel,1989).The poorly crystalline ferric arsenate synthesized in this work was determined to have the formula Fe1.02AsO4Æ2.4H2O.Fig.1shows the effect of pH on the infrared spectra of the sorption samples of arsenate on sulfate–ferrihydrite(initial molar ratio of Fe/As=2).Both detailed display of the As–O stretching vibration region(500–1000cmÀ1)and the whole range of the scanning(400–4000cmÀ1)are shown in thefigure.Poorly crystalline ferric arsenate shows a strong well-resolved band at838cmÀ1.Within the crystalline ferric arsenate(i.e.scorodite)structure,AsO4tetrahedra and FeO4(OH2)2octahedra connect alternately at vertices(Kita-hama et al.,1975).The arsenate is coordinated with four iron octahedra with an average As–O bond length of1.68A˚.The band at838cmÀ1was attributed to the stretching vibration of As–O coordinating to iron atom,i.e.As–O–Fe.The weakThe nature of adsorbed arsenate on ferrihydrite1645shoulder at$750cmÀ1was probably caused by the hydrogen bonding between H2O and AsO4since the H-bonding re-sulted in increased bond length and a red shift of the wave number(Myneni et al.,1998).The1625cmÀ1band was due to water O–H bending mode whereas the stretching vibration bands of O–H were located at$3194and$3373cmÀ1.The bands between950and1250cmÀ1were assigned to struc-tural SO42Àions,which were incorporated into the poorlycrystalline ferric arsenate by substitution of AsO43Àions dur-ing synthesis from sulfate medium.Sulfate ions were incorpo-rated into crystalline scorodite synthesized from sulfate solution(Singhania et al.,2005).The infrared spectra of pH3and4sorption samples also exhibited a strong,well-resolved band in the As–O stretching vibration region at similar position($833cmÀ1)to that of poorly crystalline ferric arsenate,indicating similarities of the arsenate bonding structures between sorption samples and poorly crystalline ferric arsenate.This suggested the for-mation of a ferric arsenate surface precipitate in the arsenate–ferrihydrite sorption samples synthesized in acidic media. However,the weak shoulder at$750cmÀ1on the FTIR spec-trum of the poorly crystalline ferric arsenate was missing for the sorption solids for some unknown reasons.The band at$833cmÀ1for the pH3and4sorption samples was assigned to As–O stretching vibration of the As–O–Fe coordination of ferric arsenate precipitate on fer-rihydrite.The formation of ferric arsenate phase in pH3 and4sorption samples was also supported by the O–H stretching vibration band at$3190cmÀ1.All samples showed a strong broad O–H band at$3370,but only theacidic sorption samples displayed the$3190cmÀ1O–H stretching vibration band like the case of poorly crystalline ferric arsenate.This characteristic O–H band of poorly crystalline ferric arsenate at$3190cmÀ1was fading out with increasing pH,indicating the disappearance of ferric arsenate surface precipitate at neutral and alkaline pH.As pH increased from3to8,the As–O stretching vibra-tion band shifted gradually from$833cmÀ1down to $806cmÀ1.At the same time,a new band emerged at high-er frequency(870–880cmÀ1)and its intensity was more pronounced with increasing pH.At pH8,we could clearly see the splitting of the single band into two bands.Peak deconvolution and curvefitting of the band produced two peaks at$806–810and$878cmÀ1(Fig.2).It is well estab-lished that at mildly alkaline pH,arsenate is adsorbed on ferrihydrite via bidentate binuclear complexation with surface iron polyhedra(Harrison and Berkheiser,1982; Waychunas et al.,1993).The band at$878cmÀ1of the pH6–8sorption samples was assigned to uncomplexed/ unprotonated As–O,whereas the$806–808cmÀ1arose from the two As–O–Fe complexed to ferrihydrite surface. Two infrared bands at824/861and817/854cmÀ1were ob-served for the arsenate adsorbed on amorphous iron oxide at pH5and9,respectively(Goldberg and Johnston,2001). The lower frequency bands at817and824cmÀ1were as-signed to the stretching vibration of As–O–Fe and the high-er frequency bands at854and861cmÀ1were attributed to ‘‘non-surface-complexed’’As–O bonds of the adsorbed arsenate species(Goldberg and Johnston,2001).It is inter-esting to note that the lower frequency band increased from 817and824cmÀ1as pH increased from5to9.This obser-vation and the band assignments are in good agreement with present work.Roddick-Lanzilotta et al.(2002)also reported that the As–O stretching vibration band of the ad-sorbed arsenate on ferrihydrite shifted from$825to $800cmÀ1as pH increased from2.6to8.In the case of bidentate binuclear complexation,two of the four As–O bonding structures are complexed to iron atoms(i.e.As–O–Fe)and the remaining two are present either both as unprotonated As–O or one as unprotonated As–O and the other one as protonated As–O–H.In com-parison,arsenate ions are coordinated to four iron atoms in ferric arsenate.According to Myneni et al.(1998),the force constant of the As–OM bond increases with coordina-tion number and decreases compared to uncomplexed As–O.Hence,for the bidentate adsorbed arsenate ion,the force constant of the two coordinated As–O–Fe is lower than that of the As–O–Fe in ferric arsenate,whereas the uncomplexed/unprotonated As–O bond has larger force constant compared to ferric arsenate.Consequently,the stretching vibration frequency of the uncomplexed/unprot-onated As–O is located at higher position while the fre-quency of the complexed As–O–Fe band is located at lower position.The increase and decrease of the As–O force constant for the bidentate binuclear complexed arsenate compared to ferric arsenate is supported by the As–O bond length(two at1.62,1.67A˚and the other two at1.71A˚, compared to1.68A˚of ferric arsenate)(Sherman and Ran-dall,2003).The shorter bond distance results in a stronger force constant and consequently higher infrared frequency.1646Y.Jia et al./Geochimica et Cosmochimica Acta71(2007)1643–1654A very weak band was observed at$700cmÀ1on the infrared spectrum of pH8arsenate–ferrihydrite sorption sample(see Fig.1).This band was reasonably assigned to protonated As–O–H bond of the adsorbed arsenate species, which was located at similar position to that of aqueous protonated arsenate species(Myneni et al.,1998).Complex-ation with metals cannot give such a low As–O stretching vibration frequency.The presence of protonated arsenate species was also proposed for the adsorption of arsenate on freshly prepared hydrous iron oxide and goethite(Myn-eni et al.,1998).It was noted that the weak As–O–H band at$700cmÀ1was absent at acidic pH indicating the ab-sence of protonated adsorbed arsenate species on ferrihy-drite at acidic pH.In a previous study using a dispersion infrared instrument(Harrison and Berkheiser,1982),three bands at875,805and700cmÀ1were observed for the ad-sorbed arsenate on hydrous ferric oxide(HFO)at pH6.5. They are very similar to the infrared bands of the pH8 sorption samples of this work(878,806and700cmÀ1).pH3and8are the extreme cases for the adsorption of arsenate on ferrihydrite in this work.At pH3,a sur-face precipitate developed and the adsorbed arsenate spe-cies were present mainly as poorly crystalline ferric arsenate.The possibility of surface precipitation of arse-nate on ferrihydrite was also suggested previously by Stanforth(1999)and Carlson et al.(2002).According to the latter research,a poorly crystalline ferric hydroxy-arsenate(FeOHAs)surface precipitate was found to form during adsorption of arsenate on schwertmannite and fer-rihydrite at pH3(Carlson et al.,2002).Similarly,surface precipitation of phosphate on goethite has been proposed in recent studies(Zhao and Stanforth,2001;Ler and Stanforth,2003).It was suggested that the adsorption reaction consisted of two phases:thefirst phase of rapid surface complexation followed by the second phase of slow buildup of a surface precipitate(Zhao and Stan-forth,2001).At the other extreme(i.e.pH8)arsenate was adsorbed via bidentate binuclear complexation with surface iron atoms in the form of unprotonated and probably protonated arsenate species as well(i.e.B FeO2AsðOÞ22Àand B FeO2AsðOÞðOHÞÀ, where B Fe represents the surface of ferrihydrite).This is in good agreement with the model proposed previously by Myneni et al.(1998),Jain et al.(1999)and Goldberg and Johnston(2001).Myneni et al.(1998)suggested that both protonated and unprotonated arsenate species were present as surface arsenate species adsorbed on ferrihydrite at alkaline pH.The surface arsenate specieswere proposed to be XH2AsO4,XHAsO4À,XAsO42À(X was Al or Fe)(Jain et al.,1999;Goldberg and Johnston, 2001).As pH increased from3to8,surface arsenate species shifted from poorly crystalline ferric arsenate precipitates to bidentate surface complexes.This is reasonable since poorly crystalline ferric arsenate is stable only at acidic pH and decomposes with increasing pH(Krause and Ettel, 1989).For the sorption systems whose media pHs lay be-tween3and8,both types of arsenate species were probably present on the surface of ferrihydrite,with the feature of poorly crystalline ferric arsenate being more pronounced at lower pH and the feature of bidentate complexes being more pronounced at higher pH.Similar to the Fe/As=2systems,the single As–O stretching vibration band shifted down gradually and split into two bands as pH increased from3to8for the Fe/As= 4sorption samples(Fig.3).The single band at acidic pH was assigned to As–O–Fe of ferric arsenate surface pre-cipitate.The presence of$3190cmÀ1O–H stretching vibra-tion band also indicated the development of ferric arsenate in the sorption solids at acidic pH.This band was absent onThe nature of adsorbed arsenate on ferrihydrite1647the infrared spectra of sorption samples at alkaline pH.The two bands(808and878cmÀ1)at alkaline pH are attributed to As–O–Fe bidentate–binuclear coordinating to ferrihy-drite and the uncomplexed/unprotonated As–O bond, respectively.The presence of protonated As–O–H bond could not be ruled out,since there appeared to be a weak feature at$700cmÀ1.It was interesting to note that all sorption samples from pH3to8had very similar sorption density(i.e.As/Fe$0.25,see Table1),but the surface arse-nate species varied from poorly crystalline ferric arsenate to bidentate–binuclear surface complexes as pH increased. This was indicative that the nature of surface arsenate spe-cies strongly depended on the pH of the sorption media.The spectra of pH3and4sorption samples showed strong sulfate bands between950and1250cmÀ1.Sulfate ions were apparently adsorbed onto the ferrihydrite during synthesis from sulfate medium and displaced by arsenate during adsorption as discussed elsewhere(Jia and Demopo-ulos,2005).At higher pH(i.e.5–8),the sulfate ions were substituted by arsenate ions as indicated by the absence of sulfate bands on the infrared spectra.For the adsorption systems of Fe/As=2,no sulfate band was observed for all pH samples(see Fig.1).Ferrihydrite synthesized from sulfate media was found to adsorb significantly more arsenate than that from nitrate media(Jia and Demopoulos,2005),which could not be ex-plained by the difference in BET surface areas.Therefore,it was of interest to compare the surface arsenate species of the two types of arsenate–ferrihydrite sorption samples. Fig.4shows the effect of pH on the infrared spectra of arsenate adsorbed on nitrate–ferrihydrite.Apparently the spectra are similar to those of arsenate adsorbed on sul-fate–ferrihydrite(Fig.1).Both are dominated by the strong As–O stretching vibration bands at700–950cmÀ1and strong O–H stretching vibration bands at3000–3500cmÀ1.In the As–O stretching vibration region(700–950cmÀ1),the infrared spectra of acidic sorption samples (pH3,4)exhibited a strong,well-resolved single band at similar position to that of poorly crystalline ferric arse-nate.This was indicative that a ferric arsenate surface precipitate formed when arsenate adsorbed on nitrate–fer-rihydrite at acidic pH,similar to As(V)adsorption onTable1Arsenic equilibrium concentration(mg/L)and sorption density(mol-As/mol-Fe)for the adsorption of arsenate on ferrihydrite synthesized from sulfate and nitrate medium at pH3–8and initial Fe/As molar ratio of2,4and8pH Fe/As=2Fe/As=4Fe/As=8Sulfate Nitrate Sulfate Nitrate Sulfate3198(0.49)318(0.46) 1.0(0.25) 1.9(0.25)<0.02(0.125) 4321(0.46)471(0.43) 1.9(0.25) 3.4(0.25)0.2(0.125) 5440(0.44)615(0.41) 3.6(0.25)14.8(0.25)0.2(0.125) 6498(0.43)788(0.38) 5.0(0.25)29.6(0.25)0.2(0.125) 7702(0.40)1070(0.32)15.8(0.25)100(0.24)0.6(0.125) 81178(0.30)1439(0.25)141(0.24)178(0.23) 2.1(0.125)The numbers in bracket are sorption densities,i.e.As/Fe molar ratio of the solids.1648Y.Jia et al./Geochimica et Cosmochimica Acta71(2007)1643–1654sulfate–ferrihydrite.As pH increased,the single As–O stretching vibration band split into two strong bands at $806–810cmÀ1and$878cmÀ1as observed for the sul-fate–ferrihydrite samples(see Fig.1)and were attributed to the As–O stretching vibration of bidentate binuclear complexed arsenate ions.The formation of a ferric arse-nate surface precipitate at acidic pH was also evidenced by the change of the$3190cmÀ1O–H band with pH.The spectrum of ferrihydrite displayed strong NO3Àbands at1250–1500cmÀ1.These bands disappeared after adsorption of arsenate indicating that nitrate ions previ-ously adsorbed during ferrihydrite synthesis were dis-placed by arsenate ions.3.2.Effect of coverage density on the nature of adsorbed arsenateThe effect of coverage density on the nature of adsorbed arsenate on ferrihydrite was evaluated.At acidic pH(i.e. pH3,4),the As–O stretching vibration region(700–950cmÀ1)was dominated by a strong well-resolved band (Fig.5).As the Fe/As molar ratio increased from2to8(i.e.the adsorption density As/Fe decreasing from0.49to0.125)for the pH3sorption systems(see Table1),the intensity of the As–O stretching vibration band decreased markedly.The As–O stretching vibration peak was located at similar position.The pH4arsenate–ferrihydrite sorptionThe nature of adsorbed arsenate on ferrihydrite1649samples with various arsenate sorption densities exhibited very similar infrared spectra to that of pH3samples.Poor-ly crystalline ferric arsenate was the major surface arsenate species.The nature of surface arsenate species was not sig-nificantly influenced by arsenate sorption density and equi-librium concentration.At pH6,the major As–O stretching vibration peak was located at$822cmÀ1for all the initial Fe/As molar ratios used,i.e.2,4and8.A visible shoulder was also emerging at $878cmÀ1on the infrared spectra.A bidentate complex of arsenate was becoming detectable in addition to the major surface species of ferric arsenate.When pH increased to8,all infrared spectra exhib-ited two bands at808and878cmÀ1irrespective of the initial Fe/As molar ratios indicating that bidentate com-plexes were the dominant surface arsenate species.The equilibrium concentration of arsenic ranged from0.028 to15.8mM for the pH8sorption systems and the sorp-tion density(As/Fe)ranged from0.125to0.31.Waych-unas et al.(1993)conducted EXAFS analysis on samples with sorption density of0.001–0.1and concluded that arsenate was sorbed via bidentate surface complexation. Goldberg and Johnston(2001)used FTIR to character-ize the arsenate–iron oxide sorption solids synthesized at pH5and9with equilibrium concentration of0.1–1.0mM and suspension density of4g/L($8g/L for the present work).The obtained results showed similar As–O band to the present study,i.e.two bands located at$817–824and$854–861cmÀ1.The infrared spectrum obtained by Roddick-Lanzilotta et al.(2002)for the sorption system pH2.6and arsenic equilibrium concen-tration of$0.5mM displayed a well-resolved band $825,which is similar to the single As–O band of the acidic samples in this work.Carlson et al.(2002)pro-posed the formation of ferric hydroxyarsenate in Fe/As=2.6–3.2sorption solids prepared from solutions of pH3.Therefore,by comparing the present study with other works,it is proposed that the pH plays the most impor-tant role in controlling the nature of the surface arsenate species sorbed on ferrihydrite,whereas other conditions (e.g.coverage density,suspension density etc.)play less important factors.The fate of sulfate in synthetic ferrihydrite before and after adsorption of arsenate can also be monitored on the infrared spectra.As indicated by the bands between950 and1250cmÀ1,more sulfate was incorporated into the fer-rihydrite at acidic pH(3,4)than neutral to alkaline pH (6,8)(Fig.5).After adsorption of arsenate at pH3,4 and Fe/As=4,8,there was still measurable amount of sul-fate remaining in the ferrihydrite,while at Fe/As=2,all sulfate has been desorbed.It was proposed that sulfate ions were adsorbed on goethite as both inner-sphere and outer-sphere surface complexes at acidic pH(Peak et al.,1999). Adsorption of arsenate on the ferrihydrite involved ligand exchange with previously adsorbed sulfate ions.With increasing pH,the adsorbed sulfate ions on ferrihydrite were more easily displaceable by arsenate.At pH6,sulfate bands were visible only on the Fe/As=8infrared spectrum while no sulfate was detectable by FTIR in the pH8sam-ples.This is in good agreement with a previous study that reported the quantitative analysis of the sorption solids (Jia and Demopoulos,2005).It was suggested that surface precipitation of phosphate and arsenate on goethite occurred almost simultaneously (Zhao and Stanforth,2001;Ler and Stanforth,2003).The process of surface precipitation may involve slow dissolution of ferrihydrite,ternary complexation of Fe3+ and subsequent precipitating of arsenate(Ler and Stan-forth,2003).The process was depicted schematically in Fig.6.It was not surprising that arsenate surface precipitate formed at such a high initial Fe/As molar ratio of8where arsenic equilibrium concentration is below the detection limit(<0.02mg/L)in this work.There possibly also existed tridentate complex structure that formed initially upon contacting of the arsenate ions with ferrihy-drite given the amorphous nature of the latter.The saturation state with respect to poorly crystalline ferric arsenate in acidic media was estimated in a previous study(Jia et al.,2006).When the p K sp(ferrihydrite)=39 (Langmuir,1997)and p K sp(ferric arsenate)=22.89 (Mahoney,2002)were taken in the calculation,the log IAPð½Fe3þ ½AsO43À was obviously lower than log K sp (ferric arsenate)for the pH3,Fe/As=4and Fe/As=8 sorption systems,indicating that the systems were undersat-urated with respect to poorly crystalline ferric arsenate.For。

Journal of Environmental Sciences2010,22(5)689–695Adsorption and desorption of Cu(II)and Pb(II)in paddy soils cultivated forvarious years in the subtropical ChinaLiang Ma1,2,Renkou Xu1,∗,Jun Jiang11.State Key Laboratory of Soil and Sustainable Agriculture,Institute of Soil Science,Chinese Academy of Sciences,Nanjing210008,China.E-mail:maliang@2.Graduate University of Chinese Academy of Sciences,Beijing100049,ChinaReceived22July2009;revised08December2009;accepted21December2009AbstractThe adsorption and desorption of Cu(II)and Pb(II)on upland red soil,and paddy soils which were originated from the upland soil and cultivated for8,15,35and85years,were investigated using the batch method.The study showed that the organic matter content and cation exchange capacity(CEC)of the soils are important factors controlling the adsorption and desorption of Cu(II)and Pb(II). The15-Year paddy soil had the highest adsorption capacity for Pb(II),followed by the35-Year paddy soil.Both the35-Year paddy soil and15-Year paddy soil adsorbed more Cu(II)than the upland soil and other paddy soils.The15-Year paddy soils exhibited the highest desorption percentage for both Cu(II)and Pb(II).These results are consistent with the trend for the CEC of the soils tested.The high soil CEC contributes not only to the adsorption of Cu(II)and Pb(II)but also to the electrostatic adsorption of the two heavy metals by the soils.Lower desorption percentages for Cu(II)(36.7%to42.2%)and Pb(II)(50.4%to57.9%)were observed for the85-Year paddy soil.The highest content of organic matter in the soil was responsible for the low desorption percentages for the two metals because the formation of the complexes between the organic matter and the metals could increase the stability of the heavy metals in the soils. Key words:adsorption and desorption;Cu(II);Pb(II);cultivation chronosequence;paddy soilDOI:10.1016/S1001-0742(09)60164-9IntroductionHeavy metals can enter soils through natural pedogenicprocesses and anthropogenic activities.The concentrationsof heavy metals releasing into soil systems by pedogenicprocesses are low and depend largely on the origin andnature of the parent materials(Adriano,2001).Applicationof metals containing pesticides,commercial fertilizers,andsewage sludge as well as air-borne pollution has led to awidespread accumulation of heavy metals in agriculturalsoils.High concentrations of heavy metals in soils canaffect soil fertility,ecosystem functions,and human healthnegatively(Lothenbach et al.,1999).The distribution ofheavy metals between the solution and the solid phases of asoil is a key issue in assessing the mobility and availabilityof heavy metals in soils(Jalali and Moharrami,2007).Thetransport of metals within a soil or to groundwater dependson the metal concentration of the solution phase(Br¨u mmeret al.,1986).Adsorption and desorption reactions on thesurfaces of soils and oxides are two important process-es controlling the concentration of heavy metals in soilsolutions(Br¨u mmer et al.,1988;Ainsworth et al.,1994;Backes et al.,1995;McLaren et al.,1998).The variabilityin adsorption affinity of heavy metals for soil surfaces*Corresponding author.E-mail:rkxu@has been attributed to the hydrolysis constant,electroneg-ative scale,Lewis acidity,charge density,and solubilityof precipitates including hydroxides and carbonates of agiven metal(McBride,1994;Pardo,2000;Sparks,2003).Therefore,the potential toxicity of heavy metals in soilsmainly depends on soil solid compositions,particularly theamount and type of clay minerals(Spark and Wells,1995),organic matter(Gerriste and Vandriel,1984),and iron andmanganese oxides(Elliott et al.,1985).The bioavailabilityand mobility of heavy metals in soils strongly dependon their physicochemical forms,i.e.,chemical fraction orspeciation(Kabata-Pendias,1993).The growth of rice needs irrigation and thus paddy soilsare easier to be polluted by heavy metals than uplandsoils.Wastewater irrigation has a long history in China,which has resulted in serious pollution of paddy soils byheavy metals in some regions of China(Li et al.,2009).For example,the Cd content in the heavily polluted soilsreaches5–7mg/kg in the Zhangshi Irrigation Area ofShenyang,China and its content in rice harvested from thesoil is1–2mg/kg(Wu et al.,1989).The Pb concentrationsin paddy soils in the Pearl River Delta,South China,are20%higher than those in upland soils(Wong et al.,2002).Soil contamination by Pb,Cu and Cd is of great concernbecause these metals are toxic to plants,human beings,and animals.Hence,the investigation of adsorption and690Liang Ma et al.V ol.22desorption of heavy metals in paddy soils is importance for understanding the fate of these toxic metals in soils. Soil organic matter can enhance or inhibit the adsorp-tion of heavy metals,depending on other soil properties or experimental conditions such as pH,cation exchange capacity(CEC),and accompanying ions(Gerritse and van Driel,1984;Basta et al.,1993;Harter and Naidu,1995; Romkens and Dolfing,1998).Previous investigations in-dicate that the removal of organic matter can decrease the adsorption of Cu(II)and Zn(II)by acid soils,and a greater effect of the organic matter on the Cu(II)adsorption than that on the Zn(II)adsorption was observed(Silveira et al.,2002;Agbenin and Olojo,2004).The organic matter content in paddy soils can increase with the cultivation time of rice(Li et al.,2005)and thus can affect the chemical behaviors of heavy metals in the soils.The content and type of iron oxides in paddy soils can also change with rice cultivation time as a result of the variation of redox conditions in the soils.This can also affect the adsorption of heavy metals in the soils.Red soils are distributed widely in the tropical and subtropical regions of the southern China.After they are used for growing rice and other semiaquatic crops,the change of the surface chemical properties of these soils as a result of the variation of redox conditions can affect the chemical behaviors of heavy metals,because heavy metal adsorption is closely related to the net negative surface charge density of the soils(Naidu et al.,1994,1998).The objectives of this study were:(1)to study the adsorption and desorption of lead and copper by the paddy soils cultivated for various periods collected in Yingtan,Jiangxi Province,China;(2)to analyze the relationship between the heavy metal adsorption and the soil surface chemical properties;and(3)to probe the adsorption mechanisms of the heavy metals in the paddy soils.1Materials and methods1.1Soil samplesThe surface horizons(0–20cm)of four paddy soils and one upland soil were collected from Yingtan,Jiangxi Province,China,in2007.The paddy soils were originated from the soil similar to the upland soil and used for rice cultivation for8,15,35or85years(named as8-Year, 15-Year,35-Year,and85-Year).The upland red soil was derived from the Quaternary red clay.The soil samples were air-dried and ground with a wooden roller and then passed through20-mesh and60-mesh sieves.Selected properties of the soils are given in Table1.The samples of the8-Year and15-Year paddy soils were collected from long-term experimentalfields at the Ecological Experimental Station of Red Soil,the Chinese Academy of Sciences located in Yingtan,Jiangxi Province, China.The paddyfields were double-cropped with appli-cation offlood irrigation and NPK fertilizers.The straw of thefirst rice was incorporated into the soil after harvesting (Li et al.,2005).The cultivation durations for the35-Year and85-Year paddy soils were confirmed by the records ofTable1Selected properties of the soils studiedSoil Organic Soil CEC Free iron Amorphous matter pH(cmol c/kg)oxides iron oxides(g/kg)(g/kg)(g/kg) Upland11.25 4.687.1634.87 1.148-Year16.50 5.378.0642.46 2.7115-Year20.17 5.4110.8229.72 2.4835-Year29.27 5.448.9030.76 1.7785-Year34.13 5.357.2015.55 1.09 CEC:cation exchange capacitythe local farmers as well as the Second National Chinese Soil Survey(Li et al.,2004).These twofields were near the long-term experimentalfields of the Experimental Station. After the15-Year paddyfield was established,similar soil management and crop rotation were adopted for the35-Year and85-Year paddyfields.Before the1960s,only farmyard manure and green manure were applied to the paddyfields,but in the1960s and the1970s,both organic manure and chemical fertilizers were applied.Since the 1980s,the amount of chemical fertilizers used in the paddy fields has been increased.The soil organic matter was measured by the K2Cr2O7 method(Pansu and Gautheyrou,2006).The soil pH was measured at the soil/water ratio of1:2.5(W/W)using a pH meter equipped with a combination electrode.The CEC of the soils was determined by the ammonium acetate method(Pansu and Gautheyrou,2006).The free iron oxides were extracted by the sodium dithionate-citrate-bicarbonate(DCB)method and the amorphous iron oxides by the acidic ammonium oxalate method(Pansu and Gautheyrou,2006).The concentration of Fe in the extracts was determined by atomic absorption spectrometry(AAS, TAS-990,Beijing Purkinje General Instrument Co.,Ltd., China).1.2Batch adsorption experimentsSolutions with various concentrations of Cu(II)and Pb(II)(0–2.5mmol/L)were prepared in0.01mol/L of KNO3with Cu(NO3)2or Pb(NO3)2.The pH of the so-lutions was adjusted to4.6with HNO3or NaOH.Soil samples of1.000g in duplicates were equilibrated with 25mL of the solutions in100-mL polyethylene centrifuge tubes for24hr at(25±2)°C in a constant-temperature shaker.Then,each of the solutions was separated from the solid phase by centrifugation at5000r/min for10 min.The lead and copper in the solutions were determined by AAS(TAS-990,Beijing Purkinje General Instrument Co.,Ltd.,China)and the amount of the heavy metals adsorbed by the soils was calculated from the difference between the total amount added and the amount remained in the equilibrium solution.The adsorption isotherms were obtained by plotting the amount of heavy metals adsorbed against their concentrations in the equilibrium solutions. To study the effect of pH on the adsorption of Cu(II),a series of the solutions containing1.0mmol/L of Cu(II)and 0.01mol/L of KNO3were prepared with various pH was from3to7.The same procedures as those for obtaining the adsorption isotherms were applied to study the adsorptionNo.5Adsorption and desorption of Cu(II)and Pb(II)in paddy soils cultivated for various years in the subtropical China691 of copper at different pH.The control tests were conductedby shaking the solutions without soil samples added.1.3Desorption of lead and copperAfter the separation of soil solid from the adsorptionequilibrium solution in the experiments for adsorptionisotherms,25mL of1.0mol/L KNO3solution was addedto replace the adsorbed heavy metals.The soil suspensionswere shaken for1hr at(25±2)°C in the constant-temperature shaker and then the solutions were separatedby centrifugation at5000r/min for10min(Xu et al.,2005).2Result and discussion2.1Adsorption isotherms of heavy metalsThe isotherms for the adsorption of Cu(II)and Pb(II)on the upland soil and paddy soils(Fig.1)show thatthe adsorption of Cu(II)and Pb(II)by the paddy soils isgreater than that by the upland soil.This suggests that ricecultivation is probably responsible for the increase in theadsorption capacity for the heavy metals by the soils.The15-Year paddy soil shows the highest adsorption capacityfor Pb(II),followed by the35-Year paddy soil.This is con-sistent with the trend for the soil CEC.As shown in Table1,the15-Year paddy soil has the greatest CEC and theCEC values for the15-Year paddy soil and35-Year paddysoil are greater than those for the other paddy soils andq e =1bX m+C eX m(1)where,C e(mmol/L)is the equilibrium concentration of the species in the aqueous solution,q e(mmol/kg soil)is the amount of adsorbed species,b is a constant related to the bonding strength,and X m(mmol/kg)is the maximum sorption.The Freundlich sorption equation is also commonly used as shown below(Eq.(2)):log q e=log k+1nlog C e(2)where,k and1/n are the Freundlich constants related to the adsorption capacity and intensity,respectively.Both Langmuir and Freundlish equations were used tofit the adsorption data.Thefitting results show that these two adsorption equations canfit the data well.The calculated parameters of the equations for various experi-ment systems are provided in Table2.All the correlation coefficients(R2)are greater than0.95.Therefore,both the Langmuir and Freundlish equations can be used to describe the adsorption of Cu(II)and Pb(II)by these soils.692Liang Ma et al.V ol.22 Table2Parameters of Langmuir and Freundlich equations for theadsorption of Cu(II)and Pb(II)by various soilsMetal Soil Langmuir constant Freundlich constantX m b R2K n R2Cu(II)Upland21.55 4.030.9816.66 2.220.998-Year29.85 4.590.9524.31 2.180.9915-Year31.15 6.420.9828.05 2.27 1.0035-Year30.6714.820.9829.26 3.340.9885-Year27.5510.390.9825.74 2.830.98Pb(II)Upland25.3810.650.9923.33 2.960.998-Year33.008.910.9732.27 2.370.9815-Year40.1613.110.9841.30 2.66 1.0035-Year33.7812.870.9935.34 2.530.9885-Year30.4011.340.9931.03 2.470.95The maximum adsorption of Cu(II)and Pb(II)by thesoils can be estimated from the Langmuir equation(Table2).The change in the trends of the maximum adsorption ofCu(II)and Pb(II)and the soil CEC with the periods of ricecultivation are presented in Fig.2.Both the maximum ad-sorption of the two metals and the soil CEC increase withthe increase in the cultivation duration and then decreasewith the further increase in the cultivation duration afterreaching the maximum value at the15th year.These resultsprovide an evidence to the interpretation that the soil CECis an important factor influencing the adsorption of Cu(II)and Pb(II)by the soils.The change in the soil CEC is mainly due to thereduction and dissolution of the iron oxides on the mineralsurface and the increase in the soil organic matter at theearly stage of the paddy soil formation.The weathering ofthe soil minerals is probably responsible for the decrease inthe soil CEC at the later stages of the paddy soil formation.Red soils widely distributed in the subtropical regionsusually carry both the positive and negative charges ontheir surfaces as a result of the presence of Fe/Al oxides inthe soils(Yu,1997).Iron oxides can coat soil minerals andthus mask their negative surface charge(Yu,1997).Duringthe transformation of upland red soils to paddy soils,thereductive dissolution of iron oxides can cause an increasein the total negative charge(Yu,1985).The increase inorganic matter of paddy soils with cultivation duration isNo.5Adsorption and desorption of Cu(II)and Pb(II)in paddy soils cultivated for various years in the subtropical China693 Pb(II)follows the order:15-Year paddy soil>35-Yearpaddy soil>8-Year paddy soil,upland soil>85-Yearpaddy soil.The amount of Pb(II)desorbed from the85-Year paddy soil is the lowest because of the highest contentof organic matter and the lowest CEC of the soil.Thedesorption percentage(P des)for the two heavy metals canbe calculated from the adsorption data(Fig.1)and thedesorption data(Fig.3)using the following Eq.(3):P des=M dM a×100%(3)where,M d(mmol/kg)is the amount of the heavy metal desorbed by KNO3,M a(mmol/kg)is the amount of the heavy metal adsorbed by the soils.The results of the desorption percentage for Cu(II)and Pb(II)(Table3)show that all the desorption percentages are less than100%, which is consistent with the previous reports(Nriagu, 1980).This suggests that the heavy metal desorption from the soils was incomplete and not fully reversible.If the adsorbed ions can be desorbed by unbuffered salts such as KNO3,they are called electrostatically adsorbed ions(Xu et al.,2005).Therefore,the Cu(II)and Pb(II) desorbed by KNO3were previously adsorbed by their electrostatic attraction to the soils,and thus the desorption percentages of the two heavy metals(Table3)represent the relative contribution of the electrostatic adsorption to the total adsorption of the heavy metals.Table3shows that the 15-Year paddy soil has the highest desorption percentage for Cu(II)(64.3%–71.1%),followed by the upland soil (61.7%–68.2%)and the35-Year and85-Year paddy soils have relatively lower Cu(II)desorption percentages,be-ing36.7%–39.6%and36.7%–42.2%,respectively.These results suggest that the Cu(II)adsorbed on the15-Year paddy soil and upland soil has a relatively higher mobility, because the ions adsorbed through electrostatic attraction is easy to desorb from soils to solutions.The decrease in the desorption percentage for Cu(II)desorption from the paddy soils with the increase in rice cultivation duration may be attributed to the increase in soil organic matter and the decrease in soil CEC.Cu(II)can form stable complexes with soil organic matter and this fraction of the adsorbed Cu(II)can not be desorbed by KCl(Guo et al.,2006).The amount of Cu(II)that complex with soil organic matter increases with the increase in the content of the organic matter of the paddy soils,which may account for the decline of the desorption percentage for previously adsorbed Cu(II).Higher desorption percentages for Pb(II)were observed Table3Desorption rate for Cu(II)and Pb(II)previously adsorbed onupland and paddy soilsMetal Equilibrium Upland8-Year15-Year35-Year85-Year concentration(mmol/L)Cu(II)0.261.747.864.336.736.70.460.550.667.439.041.00.867.151.770.839.642.21.268.252.971.139.141.7 Pb(II)0.274.764.772.669.057.90.472.061.472.876.653.30.670.559.372.077.051.70.868.957.372.475.250.4for the15-Year and the35-Year paddy soils.This is because their higher CEC contributes more electrostatic adsorption to the adsorption of Pb(II)by these soils.The similar mechanism is probably also responsible for the highest desorption percentage for Cu(II)observed for the 15-Year paddy soil.The desorption percentage for Pb(II) is lower for the85-Year paddy soil,ranging from50.4% to57.9%(Table3).This is similar to that observed with the Cu(II)systems discussed above.The higher content of organic matter and lower CEC are considered to be responsible for the lower desorption percentage for Pb(II) and Cu(II)observed for the85-Year paddy soil.This is because the increase in the organic matter can lead to an increase in the portion of the specific adsorption of Cu(II) and Pb(II)by the soil,and the fraction of the Cu(II)and Pb(II)adsorbed can not be desorbed by the unbuffered salts.Hence,the long-term rice cultivation probably may cause an accumulation of organic matter in the paddy soils and thus increase the stability of heavy metals such as Cu(II)and Pb(II)in these soils through the formation of complexes between the metals and the organic matter.A comparison for the two metals shows that the ad-sorption and desorption of Cu(II)are more sensitive to the effect of soil organic matter as compared to Pb(II)(Fig. 1and Table3).This may be attributed to the difference in the affinity of organic matter for the two metals.The stability constants(logβ)for the complexes of Cu(II)with citric acid(H4L),for example,are6.1(CuHL)and18 (CuL),and those for Pb(II)are5.2(PbHL)and12.3(PbL). Although both the Cu(II)and Pb(II)complexes have high stability,the stability constants for the complexes of Cu(II) are greater than those for Pb(II).Similar results were reported for the complexes of Cu(II)and Pb(II)with soil humic acid,and the stability constants for the Cu-humic acid complexes are also greater than those for the Pb-humic acid complexes(Pandey et al.,2000).Therefore, soil organic matter has a greater affinity for Cu(II)than that for Pb(II).2.3Effect of pH on Cu(II)adsorption by the soilsThe Cu(II)adsorption by the paddy soils and the upland soil followed the expected trend that the metal adsorption increased with increasing pH(Fig.4)with no adsorption observed below pH3.0(data not shown).The adsorption of Cu(II)increased sharply from pH4.0to pH 5.0and reached the maximum at about pH6.0.Increase pH reduces the solubility of most Cu-bearing minerals and increases the adsorption affinity of Cu(II)for iron oxides,organic matter,and other adsorptive surfaces.This increases cationic heavy metal retention on soil surfaces via adsorption,inner-sphere surface complexation,and/or precipitation,and multinuclear type reactions(McBride, 1994;Sparks,2003).The effect of pH on the Cu(II)adsorption may be explained as follows.At low pH,the hydroxyl groups on the surface of iron and aluminum oxides are protonated, which can lead to the competition of protons for the adsorption sites where Cu(II)can be specifically adsorbed. This may explain the observation of a very low uptake694Liang Ma et al.V ol.22No.5Adsorption and desorption of Cu(II)and Pb(II)in paddy soils cultivated for various years in the subtropical China695Air,&Soil Pollution,27(3-4):379–389.Gerritse R G,van Driel W,1984.The relationship between adsorption of trace metals,organic matter,and pH in temperate soils.Journal of Environmental Quality,13(2): 197–204.Guo X Y,Zhang S Z,Shan X Q,Luo L,Pei Z G,Zhu Y G et al.,2006.Characterization of Pb,Cu,and Cd adsorption on particulate organic matter in soil.Environmental Toxicology and Chemistry,25(9):2366–2373.Harter R D,Naidu R,1995.Role of metal-organic complexation in metal sorption by soils.Advances in Agronomy,55:219–263.Jalali M,Moharrami S,petitive adsorption of trace elements in calcareous soils of western Iran.Geoderma, 140(1-2):156–163.Kabata-Pendias A,1993.Behavioural properties of trace metals in soils.Applied Geochemistry,8(Suppl.2):3–9.Li D C,Li Z P,Zhang T L,2004.Contents of heavy metal elements in paddy soils cultivated for different years in a red soil region.Chinese Journal of Soil Science,35(3):336–338.Li P J,Wang X,Allinson G,Li X J,Xiong X Z,2009.Risk assessment of heavy metals in soil previously irrigated with industrial wastewater in Shenyang,China.Journal of Hazardous Materials,161(1):516–521.Li Z P,Zhang T L,Han F X,Felix-Henningsen P,2005.Changes in soil C and N contents and mineralization across a cultivation chronosequence of paddyfields in subtropical China.Pedosphere,15(5):554–562.Lothenbach B,Furrer G,Scharli H,Schulin R,1999.Immobiliza-tion of zinc and cadmium by montmorillonite compounds: Effects of aging and subsequent acidification.Environmen-tal Science&Technology,33(17):2945–2952.McLaren R G,Backes C A,Rate A W,Swift R S,1998.Cadmium and cobalt desorption kinetics from soil clays:Effect of sorption period.Soil Science Society of America Journal, 62(2):332–337.McBride M B,1994.Environmental Chemistry in Soils.Oxford University Press,Oxford.Naidu R,Bolan N S,Kookana R S,Tiller K G,1994.Ionic-strength and pH effects on the sorption of cadimium and the surface charge of soils.European Journal of Soil Science, 45(4):419–429.Naidu R,Sumner M E,Harter R D,1998.Sorption of heavy met-als in strongly weathered soils:An overvirw.Environmental Geochemistry and Health,20(1):5–9.Nriagu J O,1980.Zinc in the Environment.Wiley-Interscience, New York.Pandey A K,Pandey S P,Misra V,2000.Stability constants of metal-humic acid complexes and its role in environmen-tal detoxification.Ecotoxicology and Environmental Safty, 47(2):195–200.Pardo M T,2000.Sorption of lead,copper,zinc,and cadmium by soils:Effect of nitriloacetic acid on metal retention.Communications in Soil Science and Plant Analysis,31(1-2):31–40.Pansu M,Gautheyrou J,2006.Handbook of Soil Analysis-Mineralogical,Organic and Inorganic Methods.Springer-Verlag,Heidelberg,Berlin.Romkens P F A M,Dolfing J,1998.Effect of Ca on the solubility and molecular size distribution of DOC and Cu binding in soil solution samples.Environmental Science& Technology,32(3):363–369.Silveira M L A,Alleoni L R F,Camargo O A,Casagrande J C,2002.Copper adsorption in oxidic soils after removal of organic matter and iron munications in Soil Science and Plant Analysis,33(19-20):3581–3592. Sparks D L,2003.Environmental Soil Chemistry(2nd ed.).Academic Press,New York.Spark K M,Wells J D,1995.Characterizing trace metal adsorp-tion on kaolinite.European Journal of Soil Science,46(4): 633–640.Wong S C,Li X D,Zhang G,Qi S H,Min Y S,2002.Heavy metals in agricultural soils of the Pearl River Delta,South China.Environmental Pollution,119(1):33–44.Wu Y Y,Chen T,Zhang X X,1989.Pollution ecology of Cd in Zhangshi irrigation at Shenyang.Acta Ecologica Sinica, 9(1):21–26.Xu R K,Xiao S C,Zhao A Z,Ji G L,2005.Effect of Cr(VI) anions on adsorption and desorption behavior of Cu(II)in the colloidal systems of two authentic variable charge soils.Journal of Colloid and Interface Science,284(1):22–29. Yu T R,1985.Physical Chemistry of Paddy Soils.Science Press, Beijing.Yu T R,1997.Chemistry of Variable Charge Soils.Oxford University Press,New York.。

Materials Science and Engineering A391(2005)121–123Adsorption behaviors of heavy metal ions onto electrochemicallyoxidized activated carbonfibersSoo-Jin Park∗,Young-Mi KimAdvanced Materials Division,Korea Research Institute of Chemical Technology,P.O.Box107,Yusong,Daejeon305-600,South KoreaReceived4March2004;received in revised form17August2004;accepted27August2004AbstractIn this work,the effect of electrochemical oxidation treatments of activated carbonfibers(ACFs)was studied in heavy metal adsorption behaviors.As a result,the acidic or basic anodic treatment led to increases of Cr(VI),Cu(II),and Ni(II)adsorptions,which could be attributed to the oxygen-containing functional groups of the ACF surfaces.©2004Elsevier B.V.All rights reserved.Keywords:Activated carbonfibers;Electrochemical treatments;Surface properties;Heavy metal ions;Adsorption1.IntroductionAs compared with conventional granular or powder activated carbons,activated carbonfibers(ACFs)have been widely used as an excellent adsorbent because of their large surface area,microporous character,and high adsorption/desorption rate[1].Also,the microstructure of ACFs is developed during activation,and influenced by many factors,such as the degree of activation and the conditions used for carbonization[2].The adsorption/desorption rate of carbonaceous adsorbents is greatly depended on not only microporous structure but also surface properties[3].Gen-erally,the electrochemical oxidation treatment of carbon in the electrolytes used can produce microstructure and surface changes of the carbon surfaces.The advantage of electro-chemical oxidation treatment is obtaining a relatively large number of oxygen-containing functional groups on ACF surfaces[4].In this work,ACFs are modified by electrochemical ox-idation treatment with acidic or basic electrolyte to obtain oxygen-containing functional groups,and the effect of elec-∗Corresponding author.Tel.:+82428607234;fax:+82428614151.E-mail address:psjin@krict.re.kr(S.-J.Park).trochemical oxidation treatment on ACFs is studied in the context of heavy metal adsorption behaviors.2.ExperimentalThe pitch-based ACFs(bundle type,Kureha)were washed with deionized water and dried for24h at80◦C(untreated-ACFs).All other chemicals were purchased in analytical grade purity from Aldrich Chemical Co.and used as re-ceived.The ACFs were subjected to electrolytic reaction in the aqueous solutions of10wt.%H3PO4(A-ACFs)and NH4OH(B-ACFs),whereby negative ions were attracted to the surface of the ACFs acting as an anode,thereby modi-fying the ACF surfaces.A cathode graphite plate was also submerged in the electrolyte solution,and the conditions of the surface treatment were processed in an electro-bath at7A for10min.The treated ACFs emerging from the electrolytic cell were dried for6h at110◦C.The surface properties of ACFs were measured by Boehm’s titration.The specific sur-face area and the pore structure were evaluated from nitro-gen adsorption data at77K(Micromeritics,ASAP2010)[5].0.05g of the ACFs was placed in contact with150ml solution of20ppm concentrations of Na2CrO4·4H2O,CuCl2·2H2O, and NiCl2·6H2O.The single bottle was sealed with paraffin0921-5093/$–see front matter©2004Elsevier B.V.All rights reserved. doi:10.1016/j.msea.2004.08.074122S.-J.Park,Y.-M.Kim /Materials Science and Engineering A 391(2005)121–123Table 1Surface functionalities of ACFs by Boehm titrationOxygen-containing functional groups (meq/g)CarboxylicLactonic Phenolic Untreated-ACFs 34070A-ACFs 40230140B-ACFs48070film and then shaken.The adsorbed amount of heavy metal ions was measured by inductively coupled plasma-atomic emission spectrometer (ICP-AES,Jovin-Yvon Ultima-C).3.Results and discussionNumerous studies on surface functionalities of carbon are already described in the literature.The difference of the sur-face functionalities on basic ACFs and acidically treated ACFs are determined by Boehm’s titration and listed in Table 1.The three bases used in the titration are regarded as approximate probes of acidic surface functionalities accord-ing to the scheme NaHCO 3(carboxyl),Na 2CO 3(carboxyl and f-lactone),NaOH (carboxyl,f-lactonic,and phenolic).As seen in Table 1,various oxygen-containing functional groups,i.e.,carboxylic,lactonic,and phenolic groups are induced in the ACF surfaces by electrochemical oxidation treatment The porous textural parameters of the original and the modified samples are obtained,as listed in Table 2.The total pore volume and the micropore volume of elecrochemically treated ACFs are decreased.This is due to the increase of oxygen-containing functional groups,which are attributed to the block of the micropores.Also,the BET’s specific sur-face area is decreased by 19%for A-ACFs compared to the B-ACFs.This can be explained by that the specific surface area and the micropore volume of A-ACFs are decreased by the pore blocking of surface functional groups and by pore destroying of acidic electrolyte with these basic carbon ma-terials.Adsorption isotherms of Cr(VI),Cu(II),and Ni(II)from aqueous solutions of heavy metal ions in the time range 0–180min on the three samples of ACFs without adding any buffering reagent are shown in Figs.1and 2.Fig.1clearly shows that the initial adsorption rate of Cr(VI)ion on the ACFs increased rapidly,especially due to the molecular sieve structures of the ACFs.Also,the amount of Cr(VI)adsorbed is comparatively much larger than the amount of Cu(II)and Ni(II)adsorbed under similar conditions.This can be imag-ined that the high ionic radius of Cu(II)(0.70˚A)and Ni(II)Table 2Textural characteristics of electrochemically treated ACFsMicropore volume (cm 3g −1)Total pore volume (cm 3g −1)BET surface area (m 2g −1)Untreated-ACFs 0.787 1.0271944A-ACFs 0.2430.5601430B-ACFs0.6600.7811757Fig.1.Adsorption of Cr(VI)onto activated carbon fibers as a function of contacttime.Fig.2.Adsorption of metal ions onto activated carbon fibers as a function of contact time:(a)Cu(II)and (b)Ni(II).S.-J.Park,Y.-M.Kim/Materials Science and Engineering A391(2005)121–123123(0.69˚A)compared to that of Cr(VI)(0.52˚A)induces a quick saturation of adsorption sites because of steric over-crowding.It results in lower adsorption capacities for Cu(II) and Ni(II)than for Cr(VI)[6].The adsorption capacity for Cu(II)appears to be higher than that for Ni(II).It seems to be due to the difference of affinity between ACF surfaces and adsorbed ions.At the low pH,the dominant species of Cr(VI)in the solution are anionic species like HCrO4−, Cr2O72−,Cr4O132−,and Cr3O102−[7].Also,at the high pH,carbon behaves as an acid,whereas at the low pH values it behaves as a base[8].Thus,the chromate anions will be expected to interact more strongly with the electron accepter H+of surface functional groups of ACF surfaces.That is, there is a possibility of the existence of some active sites where negatively charged Cr(VI)can be adsorbed.At pH5,Cu(II)and Ni(II)are cationic species,such as, Cu2+or CuOH+and Ni2+or NiOH+[6].And they are suit-able to interact with negatively charged groups of ACFs.H+ ions compete with metal ions for the exchange sites in the system at pH5and the surface functional groups in aqueous solution produce H+ions,which are directed towards the liq-uid phase,leaving the carbon surface with negatively charged sites.Thus,the availability of relatively high concentration of negatively charged sites in case of the treated ACFs results in an increase in the adsorption of Cu(II)and Ni(II),as seen in Fig.2.As illustrated in Figs.1and2,the percentage adsorption of heavy metal ions increases with increasing the agitation reaction time and the adsorbent capacities decrease in the order B-ACFs>A-ACFs>untreated-ACFs,in spite of a de-creases of specific surface area,as seen in Figs.1and2.This is because at higher dose of adsorbent due to the increased oxygen-containing functional groups,more adsorption sites are available,causing higher removal of heavy metal ions. Also,A-ACFs show lower percentage adsorption of heavy metal ions than that of B-ACFs.It seems to be due to the destruction of micropores by acidic electrolyte,resulting in decreasing the specific surface area.These results can be ex-plained that the adsorption capacities of ACFs are greatly depended on not only microporous structure but also surface properties.4.ConclusionsIn this work,the oxidized ACFs were studied in the ad-sorption characteristics in terms of the microstructures and surface functional groups.In the results of XPS and BET, the specific surface area of the treated ACFs was decreased, whereas oxygen-containing functional groups of the treated ACFs were increased.As expected,the increased surface functional groups led to an increase of the adsorption of heavy metal ions.In case of A-ACFs,the adsorption of heavy metal ions was increased by increasing of oxygen-containing func-tional groups,in spite of a decrease of specific surface area by acidic electrolyte.References[1]A.Ahmadpour,D.D.Do,Carbon35(1997)1723.[2]Z.Yue,C.Mangun,J.Economy,P.Kemme,D.Cropek,S.Maloney,Environ.Sci.Technol.35(2001)2844.[3]S.J.Park,J.S.Shin,J.Colloid Interface Sci.264(2003)39.[4]C.Grogger,S.G.Fattakhov,V.V.Jouikov,M.M.Shulaeva,V.S.Reznik,Electrochim.Acta49(2004)721.[5]S.Brunauer,P.H.Emmett,E.Teller,J.Am.Chem.Soc.60(1938)309.[6]K.Kadirvelu,C.Faur-Brasquet,P.Le Cloirec,Langmuir16(2000)8404.[7]V.K.Garg,R.Gupta,R.Kumar,R.K.Gupta,Bioresour.Technol.92(2004)79.[8]A.¨Ozer,D.¨Ozer,J.Hazard.Mater.100(2003)219.。

简述吸附树脂的成孔技术简述吸附树脂的成孔技术引言：吸附树脂作为一种常见的分离材料，在多个领域都有广泛应用。

为了增强吸附树脂的吸附能力和效率，研究人员不断探索各种成孔技术。

本文将会简要介绍吸附树脂的成孔技术及其应用。

一、成孔技术的概念与分类1. 成孔技术的定义成孔技术是指通过在吸附树脂表面形成微孔或介孔结构，增加吸附表面积和扩大孔径，从而提高吸附树脂的吸附能力和效率的一种技术。

2. 成孔技术的分类成孔技术可分为物理成孔和化学成孔两种类型。

- 物理成孔：通过物理方法在吸附树脂表面形成孔洞或微孔，如机械打孔、电子束辐照等。

- 化学成孔：通过化学方法在吸附树脂表面引入反应产物，形成孔洞或介孔结构，如溶剂膨胀法、溶剂热法等。

二、常见成孔技术及其应用1. 机械打孔技术机械打孔技术是一种物理成孔方法，通过机械手段在吸附树脂表面直接形成孔洞。

这种方法成本低、操作简单，因此在吸附树脂制备中得到广泛应用。

具体步骤如下：- 选择一种适合的打孔设备，并调整打孔参数。

- 将吸附树脂放置在打孔设备上，进行打孔处理。

- 对打孔后的吸附树脂进行表征和测试，以确保所得成孔效果符合要求。

使用机械打孔技术制备的吸附树脂具有较大的孔径和表面积，适用于各类分离和吸附过程。

2. 溶剂膨胀法溶剂膨胀法是一种常见的化学成孔技术，通过溶剂的膨胀作用在吸附树脂表面形成介孔结构。

具体步骤如下：- 选择适当的溶剂，并确定溶剂浓度和温度。

- 将吸附树脂置于溶剂中浸泡，使其与溶剂充分接触。

- 通过溶剂膨胀和离子交换作用，在吸附树脂表面形成介孔结构。

溶剂膨胀法制备的吸附树脂具有较大的孔径和较高的孔隙度，适用于高效吸附和分离过程。

3. 电子束辐照技术电子束辐照技术是一种物理成孔方法，通过电子束的辐照作用在吸附树脂表面形成微孔。

具体步骤如下：- 选择适当的辐照设备和参数，并进行预处理。

- 将吸附树脂放置在辐照设备中，进行电子束的辐照处理。

- 对辐照后的吸附树脂进行表征和测试，以确保所得成孔效果符合要求。

第21卷第5期2021年5月________3反应与分离f过程工程学报V 。

丨.2INa5The Chinese Journal of Process Engineering May 2021DOI: 10.12034/j.issn. 1009-606X.220142Dynamic adsorption of low concentration gallium ion by LX-92 resin insulfuric acid systemC h aolu W E N 1, Zhenhua SU N 2, Shaopeng L I 2*, Zhibin M A 1*, Huiquan L I 2,31. Institute of Resources and Environment Engineering, Shanxi University, Taiyuan, Shanxi 030006, China2. CAS Key Laboratory of Green Process and Engineering, Institute of Process Engineering, Chinese Academy of Sciences, NationalEngineering Laboratory for Hydrometallurgical Cleaner Production Technology, Beijing 100190, China 3. College of Chemical Engineering, University of Chinese Academy of Sciences, Beijing 100049, ChinaAbstract: The dynamic adsorption -desorption behavior o f gallium in simulated sulfuric acid leach solution o f fly ash on polystyrene resin (L X -92) was investigated by fixed bed device . The dynamic adsorption process was analyzed by Thomas , Yoon -Nelson , and Adam - Bohart empirical models . The results showed that decreasing the flow rate (F a d ) andtheinitialconcentrationo fgallium (III ) (C 〇), increasing the bed height (Z ) were conducive to improve the fixed bed adsorption efficiency and equilibrium adsorptioncapacity .According to experiments data , theoptimumconditionsfordynamicadsorption process was as follow : Fa d =5.0 mL /min , C 〇=260 mg/L and r =55〇C , and the maximum adsorption capacity was 56.65 mg /g . The elution rate o f gallium could reach 94.40% at the conditions o f 3.0 mol/L H2SO4 and 1.0 mL/min flow rate . After the process o f adsorption and desorption , the concentration o f gallium ions could be enriched more than 10 times . The dynamic adsorption behavior o f gallium by the resin was well fitted by the Yoon-Nelson dynamic adsorption model . The corresponding equations o f the dynamic adsorption rate constant AV n and the half-through time r value constant with the initial Ga (III ) ion concentration , flow rate , and bed height were established . The dynamic adsorption results would be used for engineering purpose o f the o f low-concentration gallium ions recovery .Key words: gallium ; dynamic adsorption ; acid system ; low concentration收稿• 2020-04-23,修回：2020-06-23，网络发表：2020-07-31，Received: 2020-04-23, Revised: 2020-06-23, Published online: 2020-07-31基金项目：国家重点研发计划资助项目(编号：2017YFB0603102);国家自然科学基金资助项目(编呤：U1810205);山西省煤基低碳重大专项资助项目 (编号：MC2016-05)作者简介：文朝璐(丨994-)，女，山西省运城市人，硕士研究生，资源循环与科学专业，E-mail: 768丨75960@ ;通讯联系人，李少鹏，E-mail:************.cn ；马志斌，E-mail:****************.cn.引用格式：文朝璐，孙振华，李少鹏，等.L X -92树脂对硫酸体系低浓度镓离子的动态吸附.过程工程学报，2021，21(5): 567-578.Wen C L, Sun Z H, Li S N, et al. Dynamic adsorption o f low concentration gallium ion by LX-92 resin in sulfuric acid system (in Chinese). Chin. J. Process Eng., 2021,21(5): 567-578, DOI:10.12034/j.issn. 1009-606X.220142.568过程工程学报第21卷L X-92树脂对硫酸体系低浓度镓离子的动态吸附文朝璐、孙振华2，李少鹏2%马志斌'李会泉231.山西大学资源与环境T程研宄所，山西太原0300062.中国科学院过程工程研究所绿色过程与工程重点实验室，湿法冶金清洁生产技术国家工程实验室，北京1〇〇丨903.中国科学院大学化学工程学院，北京100049摘要：研究了粉煤灰模拟硫酸浸出液中的镓在聚苯乙烯树脂(LX-92)上吸附分离的可能性，采用固定床吸附装置考察了树脂动 态吸附-脱附镓的行为，利用Thomas，Yoon-Nelson和Adam-Bohart经验模.型对动态吸附过程进行f分析和预测◊结果表明，降低流速(F)、增加床层高度(Z)、减小镓(III)初始浓度(C o)有助于提高固定床吸附效率和平衡吸附容量；在C o为260 5.0 mL/min、吸附温度为55'C的条件下，树脂的最大动态平衡吸附容量为56.65m g/g:用3.〇!!1〇丨/1^5〇4在丨.〇1111_/1^11流速的最佳洗脱条件下，洗脱率达到94.40%;树脂在硫酸体系中对低浓度镓离子的吸附-脱附具有良好的循环使用性，经过吸附-脱附镓 离子可富集10倍以上；树脂吸附镓的动态行为满足Yoon-Nelson动态吸附模型，建立了动态吸附速率常数A：YN和半穿透时间r 值常数与初始离子浓度、流速、床层高度的对应方程，为低浓度镓离子的吸附法提取工程化提供了理论基础。

吸收比范围1. 引言吸收比范围是指在特定条件下，物质吸收或吸附的能力在一定浓度范围内的变化情况。

吸收比范围的研究对于理解物质在环境中的行为、优化工业过程以及开发新材料等方面具有重要意义。

本文将介绍吸收比范围的定义、影响因素以及相关应用领域。

2. 吸收比范围的定义吸收比范围是指在一定条件下，物质吸附或吸收能力随着溶液或气体浓度变化而发生变化的范围。

常用来描述物质与溶液或气体之间相互作用强度和效果的变化情况。

3. 影响因素3.1 温度温度是影响吸收比范围的重要因素之一。

一般来说，随着温度升高，物质分子运动加剧，表面活性增强，从而增加了物质与溶液或气体之间的相互作用力，使得吸附或吸收能力提高。

3.2 pH值pH值是溶液中酸碱性的度量指标。

不同物质对酸碱环境的敏感程度不同，因此pH值的变化会对吸收比范围产生影响。

一般来说，pH值偏酸性时有利于一些阳离子的吸附，而偏碱性时有利于一些阴离子的吸附。

3.3 浓度溶液或气体的浓度是影响吸收比范围的重要因素之一。

当浓度较低时，物质与溶液或气体之间相互作用力较弱，吸附或吸收能力较低；而当浓度较高时，物质分子间的相互作用力增强，吸附或吸收能力增加。

3.4 表面特性物质的表面特性也会对其吸收比范围产生影响。

例如，材料的孔隙结构、表面活性、化学成分等都会影响物质与溶液或气体之间的相互作用力和吸附效果。

4. 应用领域4.1 环境领域在环境领域，吸收比范围的研究可以帮助我们理解污染物在土壤、水体和大气中的行为。

通过研究吸收比范围，可以确定最佳处理条件，提高污染物的去除效率。

此外，吸收比范围还可以用于评估环境风险和制定环保政策。

4.2 工业领域在工业领域，吸收比范围的研究对于优化工艺过程非常重要。

通过了解物质与溶液或气体之间的相互作用力变化情况，可以选择合适的吸附材料、调节处理条件，提高产品质量和产能。

4.3 材料科学领域在材料科学领域，吸收比范围的研究有助于开发新型材料。

通过调控材料表面特性和化学成分等因素，可以实现对特定物质的高效吸附或吸收。

两性离子表面活性剂改性蒙脱石结构特征及其吸附性能的研究的开题报告一、研究背景蒙脱石是一种重要的天然无机材料，具有吸附、离子交换等多种应用价值。

然而，它的应用受到其晶体结构的限制。

因此，对蒙脱石进行改性，使其具有更好的吸附性能，具有重要的研究意义和应用前景。

二、研究内容本研究将采用两性离子表面活性剂对蒙脱石进行改性，并研究改性后的蒙脱石的结构特征及吸附性能。

具体研究内容如下：1. 蒙脱石的结构特征分析，包括X射线衍射、扫描电镜等技术；2. 采用不同浓度、不同类型的两性离子表面活性剂对蒙脱石进行改性，并通过吸附等温线等方法研究改性后的蒙脱石的活性表面积及吸附性能；3. 系统研究改性蒙脱石的吸附性能对离子浓度、pH值等因素的影响；4. 利用文献资料和理论分析，探究两性离子表面活性剂对蒙脱石表面改性的机理。

三、研究意义1. 本研究可以为改进蒙脱石吸附性能提供参考，促进其在环境治理、废水处理等领域的应用；2. 探究两性离子表面活性剂对蒙脱石表面改性的机理，对于理解表面改性过程具有一定的指导意义。

四、研究方法实验室制备蒙脱石样品，采用X射线衍射、扫描电镜等技术对其结构特征进行表征；采用吸附等温线、差示扫描量热等技术研究改性后的蒙脱石的活性表面积及吸附性能；利用变量汇聚法进行吸附性能与离子浓度、pH值等因素的关联分析；通过文献资料和理论分析，探究两性离子表面活性剂对蒙脱石表面改性的机理。

五、预期成果1. 探究两性离子表面活性剂对蒙脱石表面改性的机理；2. 揭示两性离子表面活性剂改性对蒙脱石结构特征及吸附性能的影响；3. 形成一定的理论研究结果和实际应用价值。

六、研究进度计划第一年：实验室制备蒙脱石样品，并进行结构表征；初步研究两性离子表面活性剂对蒙脱石表面改性的影响；第二年：系统研究两性离子表面活性剂改性对蒙脱石的吸附性能的影响；第三年：深入探究两性离子表面活性剂改性对蒙脱石表面的机理；并总结实验和理论研究成果，形成一定的研究报告。