英文翻译 附原文

本科毕业设计(论文) 外文翻译（附外文原文）

系 ( 院 )：资源与环境工程系

课题名称：英文翻译

专业(方向)：环境工程

班级：2004-1班

学生：3040106119

指导教师：刘辉利副教授

日期：2008年4月20

使用褐煤（一种低成本吸附剂）从酸性矿物废水中去除和回收金属离子

a. 美国, 大学公园, PA 16802, 宾夕法尼亚州立大学, 能源部和Geo 环境工程学.

b. 印度第80号邮箱, Mahatma Gandhi ?Marg, Lucknow 226001, 工业毒素学研究中心, 环境化学分部,

于2006 年5月6 日网上获得，2006 年4月24 日接受，2006 年3月19 日;

校正，2006 年2月15 日接收。

摘要

酸性矿物废水(AMD), 是一个长期的重大环境问题，起因于钢硫铁矿的微生物在水和空气氧化作用, 买得起包含毒性金属离子的一种酸性解答。这项研究的主要宗旨是通过使用褐煤（一种低成本吸附剂）从酸性矿水(AMD)中去除和回收金属离子。褐煤已被用于酸性矿水排水AMD 的处理。经研究其能吸附亚铁, 铁, 锰、锌和钙在multi-component 含水系统中。研究通过在不同的酸碱度里进行以找出最适宜的酸碱度。模拟工业条件进行酸性矿物废水处理, 所有研究被进行通过单一的并且设定多专栏流动模式。空的床接触时间(EBCT) 模型被使用为了使吸附剂用量减到最小。金属离子的回收并且吸附剂的再生成功地达到了使用0.1 M 硝酸不用分解塔器。

关键词:吸附; 重金属; 吸附; 褐煤; 酸性矿物废水处理; 固体废料再利用; 亚铁; 铁; 锰。

文章概述

1. 介绍

2. 材料和方法

2.1. 化学制品、材料和设备

3. 吸附步骤

3.1. 酸碱度最佳化

3.2. 固定床研究

3.2.1 单一栏

3.2.2 多栏

4. 结果和讨论

4.1. ZPC 和渗析特征

4.2 酸碱度的影响

4.3. Multi-component 固定吸附床

4.3.1 褐煤使用率

4.4. 吸附机制

4.5. 解吸附作用研究

5. 结论

1. 介绍

酸性矿物废水(AMD) 是一个严重的环境问题起因于硫化物矿物风化, 譬如硫铁矿(FeS2) 和它的同素异形体矿物(α-FeS) 。它表征为一种低酸碱度和高浓度硫化物和金属

[ 1 ] 和[ 2 ] 。AMD 通常包含高浓度的金属譬如钢、锰、锌和少量镉、铅、铜, 和镍。硫化物的氧化作用而溶化释放的亚铁钢和酸度进入河水随后释放其它金属离子。AMD 典型地包含高浓度和溶解性离子可能存在高浓度(Fe2+) 或以被氧化的形式(Fe3+) 。酸度在AMD 包括氢离子酸度和矿物酸度(钢、铝、锰, 和其它金属离子依赖特别的地质环境和金属硫化物) 。在美国有20,000-50,000 个被摒弃的矿。许多这些矿产品导致酸性矿物废水(AMD) 。这AMD 包含各种各样的重金属, 有害地影响23,000 公里小河在美国[ 3 ] 并且[ 4 ] 。

宾夕法尼亚州单独有25% 所有被摒弃的矿站点由引起酸性废水表层开采的办公室列出。超过3000 英里的河在宾夕法尼亚州单独被酸性矿物废水污染在这个国家。如果留下未经处理, 酸性矿物废水可能污染地表和地面水, 损坏生态环境，野生生物，和鱼类[ 5 ] 。当酸性矿物废水与水面混合有潜在的为总污染。茶黄(氧化铁) 沉淀物可能覆盖水资源和杀害水生植物群。生物活动损失能摧残食物链和最后导致鱼被杀害和礼节[ 6 ] 。

钢存在在铁(Fe3+) 或亚铁(Fe2+) 形式, 取决于酸碱度, 和被溶化的氧浓度。在中立酸碱度和在氧气共同存在, 可溶解Fe2+ 被氧化的Fe3+, 水解对铁氢氧化是不能溶解在水中。在大多表面水中, Fe3+ 占优势。亚铁(Fe2+) 另一方面是可溶解和受厌氧支配。锰存在在

7+, 6+, 5+, 4+, 3+, 2+, 0 氧化状态下。最共同的氧化作用状态在水溶液是Mn(II) 。既使当MnO2 (pyrolucite) 是锰在自然里的最重要的形式, 它是可溶解在非常低酸碱度作为Mn(II): MnO2 + 2HCl = Mn2+ + Cl2 + H2O (1)

或碱性情况, 给等分子的总量(dismutation) Mn(III) 和M(V) 。锰的工业来源是钢合金, 干电池, 玻璃和陶瓷, 油漆、墨水、染料和肥料[ 7 ] 。它同时发生在金属矿, 矿物废水，特别是在煤田。对锰的主要关注于饮用水里它厌恶的口味和它的污染容量。饮用水指南由美国供水系统协会推荐(AWWA) 是0.05 mg/l 。最有效的可利用的技术(棒) 放电极为酸性矿物废水(mg/l) 被提供在表1 [ 8 ] 。

表1 。

最有效的可利用的技术(棒) 放电极为酸性矿物废水(mg/l) [ 8 ]

化作用和水解[ 12 ], [ 13 ], [ 14 ] 和[ 15 ], 中和[ 16 ] 和[ 17 ], 离子交换和溶剂提取[ 18 ], [ 19 ] 和[ 20 ], 离子交换和沉淀[ 21 ] 和[ 22 ], 滴定法[ 23 ], 生物吸附[ 24 ], [ 25 ], [ 26 ] 和[ 27 ], 吸附[ 1 ], [ 28 ], [ 29 ], [ 30 ], [ 31 ], [ 32 ], [ 33 ], [ 34 ], [ 35 ], [ 36 ] 和[ 37 ], 反渗透作用[ 38 ] 。其它用于处理酸性矿物废水的方法也被回顾[ 6 ], [ 39 ] 和[ 40 ] 。吸附作用应经演变作为去除金属离子的方法。由于活性炭用于水处理的费用较高, 寻求一种代替品正在进行中。这样的吸附剂应该是有效的, 经济上可行的, 也较容易再生。已经有研究者在寻求更廉价的材料为去除Fe(II) 、Fe(III) 、Mn(II) 从水中譬如泡沸石[ 41 ] 和[ 42 ], 被激活

的烂泥[ 43 ], 胶磷矿[ 44 ], 含水锰二氧化物[ 45 ], 和钢[ 46 ] 。用这种方法制作的其它努力被描述以各种各样的评论文章[ 47 ] 和[ 48 ] 。

对褐煤的用途在废水处理近几年来受到了增长的注意[ 49 ], [ 50 ] 和[ 51 ] 。除此之外，低廉褐煤拥有它有效的媒介为去除含水金属的几个特征。褐煤拥有高氧含量, 固定的羟基和氢氧根基团。这些基团是离子交换的活跃中心。这些褐煤材料可能被使用作为供选择的正离子交换器[ 52 ], [ 53 ] 和[ 54 ] 。羟基和氢氧根基团能在离子交换反应参与以以下方式(Eqs.

(2), (3) 和(4))

n L–COOH + M n+ = n L–COOM + n H+ (2)

n L–COOH + M n+ = (n? 1)L–COOM(OH) (3)

形成hydroxilated 种类也可能被考虑

M n+ + H2O = M(OH)n + n H+ (3)

这些反应导致了在吸附以后酸化的解答。离解常数(pKa) 羟基基团一般是在4 和6 之间。因而提高[ H+ ] 的解答归结于Eq .(1) 。对这材料的用途在去除金属离子酸性矿物废水正被调查并且结果被提出在这篇文章里。本研究的目标是为了使那些不受酸性矿物废水[ Fe(II) 和Fe(III) ], 锰, 钙和铝污染。

2. 材料和方法

2.1. 化学药品、材料和设备

材料和设备褐煤样品被使用在这项研究被获得在德克萨斯州，Martin 湖。在吸附被研究之前样品搽粉了对负号325 B.S.S 滤网。化学制品是AR 品级。测试试剂的储蓄解答被做了在被去离子的水中由Millipipore-Q? 。酸碱度被调整了使用稀释硫酸和氢氧化钠。酸碱度测量被执行了使用美国酸碱度米模型710, Orion。金属含量在样品是使用引人体结合的血浆(ICP) 分光光度表(Leeman PS 3000UV, 美国) 。碳、氢和氮气(CHN) 分析使用LECO CHN 分析仪, 并且XRD 线外形分析使用Rigaku Geigerflex X-射线衍射计。

3. 吸附机制

3.1. 酸碱度最佳化

Fe(II) 、Fe(III) 、Mn(II) 和Ca(II) 的去除在另外酸碱度批量处理方式被学习了。最初浓度的100 ppm 为Fe(II) 和Mn(II) 和60 ppm 的Fe(III) 被使用了。一25 机器固定的浓度测试被解答以0.15 g褐煤和间歇性鼓动48 h 。接触时间和条件根据初步实验被选择, 显示出, 平衡被建立在48 h之后。在这个期间被过滤了使用第42 号纸板和使用ICP分析。金属含量被保留在吸附剂阶段(qe, mg/g) 由计算公式得Eq .(5)

(5)

C0 和Ce 是金属离子的最初和最后的(平衡) 在溶液中(m), V 溶液容量(l) 和m大量的褐煤(g)。

3.2. 固定的床研究

确定褐煤吸附容量的条件, 专栏实验下被进行了以流动的方式。单一或者多专栏塔器被使用辨认各种各样的设计参数。

专栏被设定了基于以下步骤。

3.2.1. 单一专栏

大约10 g (干燥) (20 □40 滤网) 褐煤被做了成为热的泥浆用于去除水中的细小颗粒。泥浆慢慢地然后被驯化入专栏,玻璃棉被使用了作为专栏支持。固定床专栏流动方式操作。吸

附床长度是近似20 cm 和床容量是22ml。使用泵是水力流速保持在2.5 ml/min。流出样品每15 分钟采集一次, 测量酸度和酸度使用硫酸(0.2 N)的化学分析方法。专栏用金属离子解决被装载了包含Fe(II) 、Mn(II) 、Zn(II), 和Ca((II) [的混合物比例= 60:50:50:50 ppm ] 。

3.2.2. 多栏式

多专栏设定的图示被提出在图1 。专栏被连接了在系列的流动方式中。大约150 g (50-g 在各个专栏, 干燥) 褐煤是重量(20 □40 滤网) 并且做成热的泥浆用于去除水中的细小颗粒。做法被描述为填装单一专栏的重复测试。金属溶液被灌输到第一专栏。各个专栏的吸附床的长度大约16 cm 并且各个专栏的床容量是80 ml(共计240 ml) 。使用泵是水力流速保持在10 ml/min,主要为导电线速度控制器。流出样品从第三个专栏每20 或50 分钟被收集一次, 测量酸度和酸度使用硫酸(0.2 N) 和存放在40 ml试管里分析。

图1. 褐煤固定床吸附剂(三张床系统) 的设计被使用做酸性矿物废水（AMD）处理。

4. 结果和讨论

褐煤是一种固体燃料用于家庭里归类为棕色煤。宏观地, 当潮湿时褐煤是黑褐色或黑色或者当干燥时是浅褐色。它的密度在1.0-1.35 g/cm3范围内。褐煤通常是无定形和纤维状或木质的纹理。褐煤拥有高表面, 典型地100-200 m2/g [ 55 ] 和[ 56 ] 。

褐煤变成灰烬成有机物之前被确定为无机组分。灰酸化后用放射光谱仪进行消化和分析。对褐煤的化学分析显示SiO2 =32.1%; Al2.O3 =13.4%; TiO2 =1.30%; CaO =17.2%; MgO

=1.81%; MnO =1.81%; Fe2.O3=18.8%; Na2.O =0.28%; K2.O =0.23%; 钡=0.14%; SrO

=0.26%; C =55.56%; H =4.83%; N =1.14%; S =0.84% [ 2 ] 。

X-射线对样品的衍射分析表明, 褐煤主要由无定形的组织与一些石英组成。

4.1. ZPC 和浸出特征

不同数量的褐煤(2, 4, 8, 16, 32 mg/l) 增加进废水有不同的最初酸碱度(2-10)。空白没有的褐煤并且跑进去了样品。所有实验一式两份获得。试管间歇性地鼓动48 h 在室温条件下, 然后在复写纸记下。最后溶液酸碱度被测量。酸碱度平衡在不同的褐煤数量的条件下被显示

在下图2 。平衡酸碱度non-linearly 增加的酸碱度范围在2-5 。增量显著在酸性媒介条件即从2.0 到4.0之间。在更高的最初酸碱度没有增加在酸碱度[ 2 ]的溶液中。pH PZC被确定了作为没有改变在联络以后与样品水的酸碱度。根据这项研究, 褐煤pH pH PZC [ 2 ] 。

图2. 相当数量的作用吸附在酸碱度平衡时溶液中没有金属离子[ 2 ] 。

由于吸附质被广泛使用, 系统中酸度值变化必须归结于褐煤在溶液中的交换互作用L2Ca + 2H+ = L2H2 + Ca2+ (6)

L2Mg + 2H+ = L2H2 + Mg2+ (7)

这些反应可能发生从左到右在强烈酸性媒介由于氢离子的高浓度。结果钙, 镁、钠、铝和钾离子被释放入解答(图3) 。它是显然的从图3, 浸出这些金属与褐煤的数额直接地是比例(除了Al3+, Fe2+) 当前在水中在变化PH 值[ 2 ] 。变化在Al3+ 的集中上, Fe2+ 更加复杂归结于这些离子[ 2 的] 同时浸出和吸附着。

图3. 褐煤的浸出行为[ 2 ] 。

4.2. 酸碱度的影响

Fe(II) 和Mn(II)被去除在不同的PH 值被显示在图4(a) (c) 。Fe(II) 、Mn(II) 和Fe(III) 的吸着在褐煤增加以在酸碱度的增量。通过吸附研究, 这些酸碱度比酸碱度必须是较少于沉淀的金属离子。Fe(II) 的吸附是非常降低在pH<2 它增加从6% 到84% (图4(a)) pH>4。去除发生由于沉淀即氢氧根离子从溶液形成了一些复合体与Fe(II) 。并且在酸范围pHfin 增加以增长的pHin, 即中和吸附是一个平衡的过程。最宜的最初酸碱度被选择为Fe(II) 是3.5, 关联去除以吸附过程。Mn(II) 的吸附在另外最初酸碱度被显示在图4(b) 。与Fe(II) 相似去除是在低pHin<2和它增加以增量在pHin 和在pHin>8.0 沉淀发生在[2 ] 。所以最宜酸碱度被选择为6.0 。并且pHfin 增加在更加酸性的pHin 并且在pHin 4.0 以后, pHfin 变得更或较不恒定。这些离子较少去除在低酸碱度归因于在这项研究中氢核和金属离子。Fe(III) 去除在另外最初酸碱度被显示在图4(c) 。吸附是与其它二种金属离子不同。那里在吸附过程的增量是在酸碱度为2.0, 和pHin >3.0 发生吸附和沉淀在[ 2 ] 。平衡酸碱度增加的pHin 范围在2-3 。最宜的最初酸碱度被选择为平衡研究是2.7-2.8 关联去除的吸附过程。进一步, 在酸碱度的减退作为金属含量功能可能归结于二平衡:

离子交换反应: n LH + M n+ = n LM + n H+ (8)

水解反应：M n+ + n H2O = Mn(OH)n + n H+ (9)

Fe(II) 在水溶液中的化学作用并不明显。在水中, 它迅速地氧化Fe(III) (E0 Fe(III)/Fe(II) =

0.771 V) 。Fe(III)被显示是一个酸性正离子, 和倾向于沉淀作为Fe(OH)3 (pKps = 35) 酸度

强烈的减少显示在图4(c) 。

图4. 酸碱度对(a) Fe(II)的吸附作用，(b) Mn(II) (100 mg/l 和褐煤用量6 g/l), 和(c) Fe(III) (50 mg/l 和褐煤用量6 g/l) 在褐煤[ 2 ] 。

4.3. Multi-component 固定吸附床

在酸性矿物废水除了存在除钢和锰以外，因此褐煤合适其它金属离子的吸附作用。multi-component 系统是复杂的由于溶质溶质竞争和溶质表面互吸附作用。

Multi-component 互作用发生在固液平衡阶段将涌现显示唯一金属离子另外容量的活跃吸附点。multi-component 系统的解释证明是复杂的, 也许是作用的一个或所有以下参量: 离子半径、负电性、酸碱度和活跃站点的可及性在吸附剂。

这项研究测试了这二种类型专栏: (a) 单一专栏测试和(b) 多个专栏测试。专栏测试步骤的具体做法被给出在第二部分。sorbable 杂质被去除当液体通过吸附床, 并且吸附发生在床的部份被定义作为质量传递吸附区域或MTZ 。

突破曲线为唯一和三个专栏系统被提出在图5 和图6Figs 。5(a) 和6(a), 各自地根据无维的浓度速率(Ce/C0) 对流出容量。并且变化在酸碱度上以流出容量被提出在图5 和图6 。在唯一专栏吸附行为被发现相当不同的比那三个专栏系统由于在空的床接触时间上的区别。在唯一专栏测试, 空的床接触时间是低于

(9 分钟), 所有金属离子的吸附被发现相对地小于Fe(II) 的最高的吸附在他们之中并且突破迅速发生。吸附作用顺序Fe(II)>Zn(II)>Mn(II)

Ca2+ + Zn2+) 唯一专栏吸附条件在25 .C 和酸碱度3.5 。

图6. 固定床(a) 突破和(b) 污水酸碱度曲线为multi-component 金属离子(Fe2+ + Mn2+ + Ca2+ + Zn2+) 吸附在25 .C 和酸碱度3.5 在三个专栏系统。

在三专栏测试, 空床接触时间是24 分钟(表2) ` 超越',被保留的Fe(II) 高度偏移了早先被吸附的Mn(II) 和Zn(II) 的地方发生。这种现象共同地被观察有机化合物的吸附。因而它也许结束, 褐煤最初吸附Mn(II) 和Zn(II) 可能被放出作为Fe(II) 的液体浓度与褐煤的增量联系。在pHfin 上的变化以流出容量被提出在图5 和图6 为单一和三个专栏系统。它各自地被注意, 流出物酸碱度减少从7.5 和6.5到4.7 和4.0, 。因而它可能结束, 更多的吸附将是更高的pH fin和反之亦然。这些结果与批结果是一致的。

表2

褐煤专栏表现为金属离子的多组分吸附

空的床接触时光或停留时间。

专栏容量确定了总面积对加入流出物和划分这价值由吸附重量在专栏点。专栏的表现根据容量、床容量、突破容量等被提出在表2 。高容量在三专栏系统的情况下归结于事实空的床接触时间是高的与单一专栏系统比较。

4.3.1. 褐煤使用率

褐煤使用率(LUR) 被定义作为参量相似与碳用法率(CUR)，确定褐煤会被用尽的率并且多频繁褐煤必须被改变或再生。LUR 也许存在以下联系:

1000)1000/(?=）

碳的使用率（）每专栏褐煤重量（褐煤使用率gal lb gal lb ）

碳的使用率（）每专栏褐煤重量（褐煤使用率l g l g =)/( 使用率在lb/gal 或g/l 为两个系统被提出在表2 。它注意到, 与三个专栏系统比较，单一专栏使用率是非常高的。线概念可能被优选使用基本的使设计达到

最低成本。经营的线接近一个极小值在两个轴在图7 。它可能被注意, 在使用率最初的迅速减退以EBCT 的增量。因而从这数据EBCT 被使用为承包商的设计也许采取EBCT 以最低的碳使用率。要求增加EBCT 的好处是使经济评估处理设计得到优化(图7) 。

图7. EBRT 运行的操作界线。

4.4. 吸着机制

离子交换最主要的机制是Fe2+ 、Mn2+ 、Fe3+ 和Zn2+ 的吸附在褐煤。它被观察, 在Fe2+ 、Mn2+ 、Fe3+, 和Zn2+ 的吸附期间, 钙离子从褐煤主要被发布了在单一, 二专栏和三部分组成的系统[ 2 ] 。除Ca(II) 以外其它轻的金属离子譬如Al3+ 、Fe2+ 、K+ 、Mg2+, 和Na+ 从褐煤被发布了无意义数额。这离子交换离子的序列和量化被发布被学习了在Fe2+ 的习惯期间, Mn2+ 在褐煤。结果显示了Ca2+ 重大发行在Fe2+ 期间。它很好被提供在溶液里, 纯净的离子交换发生以等摩尔浓度即金属l(s) 区域的比与褐煤并且发布应

该是相等的。在我们的情况金属区域的价值(Fe2+ + 1/2H+) 对褐煤表面和发布(Ca2+ +

Mg2+ + 1/2Na+) 是相等的到1 以低浓度表明离子交换机制[ 2 ] 。但在更高的浓度金属区域对褐煤表面和发布是大于一个即相当数量钙被发布是较少比相当数量金属区域表明除离子交换以外其它可能性譬如沉淀, 物理吸附无法排除。

解吸附作用测试以数量被吸附的Fe(II) 和Mn(II) 被放出的离子的水溶液。这更加进一步证实, Fe(II) 、Mn(II) 和Fe(III) 在褐煤的吸附是主要化工的在自然里。实际上，一些Fe/Mn 的吸附由其它过程依照以上也许能提高褐煤对整体吸附容量的贡献。

4.5. 解吸附作用研究

当吸附用尽或当流出物从吸附床到达最大容量放电水平, 吸附剂材料的补救以及吸附剂的再生变得相当必要。燃尽的吸附剂专栏的再生是相当一个重要过程，在废水处理并且碳专栏, 总之, 服从热量再生。在这处理10-20% 吸附中由损耗通常丢失在各个周期期间并且被吸附物的补救并且不是可能的。金属离子的解吸附作用在三个专栏系统相同流速的情况下, 长度,尝试了与0.1 N 硝酸反应并且结果被提出在图8(a) (c) 。它注意到, 金属离子的回收几乎是100% 。

图8. (a) Fe2+ 回收, (b) Mn2+, 和(c) Ca2+ 从褐煤被装载的专栏使用0.1 N 硝酸在

25 .C 。

5. 结论

研究显示, 褐煤可能被使用作为金属离子吸附剂为处理被Fe(II) 、Fe(III) 和Mn(II) 污染的废水在单一和多组分系统。通过研究不同的酸碱度条件下发现最宜的酸碱度。模仿工业条件处理酸性矿物废水, 所有研究使用单一和多专栏设置的流动方式下进行。空床的接触时间(EBCT) 模型是为使吸附剂用来减到最小。金属离子的解吸附作用在三个专栏系统的流速下, 专栏的长度的相同情况达到了与0.1 N 硝酸。它注意到, 金属离子的回收几乎是100% 。解吸附作用回收将进一步减少处理系统的费用。褐煤的用途作为一个增值吸附剂对它的燃料价值应该现在被考虑。很多低成本吸附剂越来越多的应用到水处理当中。褐煤是廉价的, 多数是无毒和可利用的。褐煤的费用是近似地$25/metric 吨

(https://www.360docs.net/doc/971400999.html,/facts/2005/CarolineGeorges.shtml) 比商业上可利用的可再生的碳费用$40/lb 的便宜，(可再生的碳6-14 滤网便宜1LB; Fisher 品牌$40.33)

(https://https://www.360docs.net/doc/971400999.html,) 。另外, 褐煤的吸附容量在处理酸性矿物废水方面优于碳

[ 2 ] 。这清楚地展示了褐煤作为吸附剂在去除酸性矿物废水和回收金属离子方面的好处。固定床的吸附参数的研究在此报告中可以得到体现，现在固定床反应器获得大规模的发展和应用，高吸附容量用于去除金属离子。

翻译原文

Removal and recovery of metal ions from acid mine drainage using lignite—A low cost sorbent

a Department of Energy and Geo-Environmental Engineering, The Pennsylvania State University, University Park, PA 16802, USA

b Environmental Chemistry Division, Industrial Toxicology Research Centre, Post Box No. 80, Mahatma Gandhi Marg, Lucknow 226001, India

Received 15 February 2006; revised 19 March 2006; accepted 24 April 2006. Available online 6 May 2006.

Abstract

Acid mine drainage (AMD), has long been a significant environmental problem resulting from the microbial oxidation of iron pyrite in presence of water and air, affording an acidic solution that contains toxic metal ions.The main objective of this study was to remove and recover metal ions from acid mine drainage (AMD) by using lignite, a low cost sorbent. Lignite has been characterized and used for the AMD treatment. Sorption of ferrous, ferric, manganese, zinc and calcium in multi-component aqueous systems was investigated. Studies were performed at different pH to find optimum pH. To simulate industrial conditions for acid mine wastewater treatment, all the studies were performed using single and multi-columns setup in down flow mode. The empty bed contact time (EBCT) model was used for minimizing the sorbent usage. Recovery of the metal ions as well as regeneration of sorbent was achieved successfully using 0.1 M nitric acid without dismantling the columns.

Keywords: Adsorption; Heavy metals; Adsorbent; Lignite; Acid mine drainage treatment; Solid waste utilization; Ferrous; Ferric; Manganese

Article Outline

1. Introduction

2. Materials and methods

2.1. Chemicals, materials and equipment

3. Sorption procedure

3.1. pH optimization

3.2. Fixed bed studies

3.2.1. Single column

3.2.2. Multi-columns

4. Results and discussions

4.1. ZPC and leaching characteristics

4.2. Effect of pH

4.3. Multi-component fixed bed sorption

4.3.1. Lignite usage rate

4.4. Sorption mechanism

4.5. Desorption studies

5. Conclusions

1. Introduction

Acid mine drainage (AMD) is a serious environmental problem resulting from the weathering of sulfide minerals, such as pyrite (FeS2) and its polymorph marcasite (α-FeS). It is characterized by a low pH-value and high levels of sulfate and metals [1] and [2]. AMD usually contains high concentration of metals such as iron, manganese, zinc and smaller amount of cadmium, lead, copper, and nickel. The oxidation of sulfide releases dissolved ferrous iron and acidity into water which subsequently releases other metal ions. AMD's typically contain high concentration of dissolved iron which may exists either in reduced form (Fe2+) or in the oxidized form (Fe3+). Acidity in AMD is comprised of hydrogen ion acidity and mineral acidity (iron, aluminum, manganese, and other metal ions depending on the specific geologic setting and metal sulfide). There are 20,000–50,000 abandoned mines in the United States. Many of these mines produce acid mine drainage (AMD). These AMD contains a variety of heavy metals, which adversely affect over 23,000 km of streams in the United States [3] and [4]. Pennsylvania alone has over 25% of all abandoned mine sites listed by the office of Surface Mining that generates mine drainage. More than 3000 miles of rivers in Pennsylvania alone are polluted with mine drainage in the nation. If left untreated, the acid drainage can contaminate surface and ground water,

damaging the health of plants, wild life, and fish [5]. When acid mine waters mix with surface waters there is potential for gross pollution. An ochre (ferric oxide) precipitate can blanket the receiving water source and kill aquatic flora and funa. The loss of biological activity could in some instances devastate the food chain and lead ultimately to fish kills and loss of amenity [6].

Iron exists in the ferric (Fe3+) or ferrous (Fe2+) form, depending upon the pH, and dissolved oxygen concentration. At neutral pH and in presence of oxygen, soluble Fe2+ is oxidized to Fe3+, which readily hydrolyzes to ferric hydroxide that is insoluble in water. In most of the surface waters, Fe3+predominates. Ferrous (Fe2+) on the other hand is soluble and dominates under anaerobic conditions.Manganese exists in 7+, 6+, 5+, 4+, 3+, 2+, 0 oxidation states. The most common oxidation state in aqueous solution is Mn(II). Even when MnO2 (pyrolucite) is the most important form of Mn in nature, it is soluble at very low pH as Mn(II):

MnO2 + 2HCl = Mn2+ + Cl2 + H2O(1) Or very alkaline conditions, to give a equimolecular amounts (dismutation) of Mn(III) and M(V). The industrial sources of manganese are steel alloy, dry cell battery, glass and ceramic, paint, ink, dye and fertilizers [7]. It also occurs in metal mines, mine drainage, especially in coal fields. The primary concerns about manganese in drinking water are its objectionable taste and its capacity to stain. The drinking water guideline recommended by American Water Works Association (AWW A) is 0.05 mg/l. The best available technology (BAT) discharge limits for acid mine wastewater (mg/l) is provided in Table 1[8].

Table 1.

Best available technology (BAT) discharge limits for acid mine wastewater (mg/l) [8]

A number of methods have been used for acid mine drainage treatment including precipitation [4], [9] and [10], electrochemical remediation [11], oxidation and hydrolysis [12], [13], [14] and [15], neutralization [16] and [17], ion exchange and solvent extraction [18], [19]and [20], ion exchange and precipitation [21]and [22], titration [23], biosorption [24], [25], [26]and [27], adsorption [1], [28], [29], [30], [31], [32], [33], [34], [35], [36]and [37], reverse osmosis [38]. Other methods used for the treatment of acid mine wastewater are also reviewed [6], [39] and [40]. Adsorption has evolved as the preferred method for metal ions removal. Due to the high cost of activated carbon for water treatment, a search for substitutes is underway. Such adsorbents should be readily available, economically feasible, and should be regenerated with ease.

Several investigators have studied less expensive materials for the removal of Fe(II), Fe(III), Mn(II) from water such as zeolites [41]and [42], activated sludge [43], collophane [44], hydrous manganese dioxide [45], and zero valent iron [46]. Other efforts made in this direction are described in various review articles [47] and [48].

The use of lignite in wastewater treatment has received increasing attention over the last several years [49], [50] and [51]. Besides being plentiful, inexpensive lignite possesses several characteristics that make it effective media for the removal of aqueous metals from wastewater. Lignites possess a high oxygen content, which is fixed in carboxyl and hydroxyl groups. These groups are infect the active centers of the ion exchange. So the lignite materials can be used as alternative cation exchangers [52], [53]and [54]. Carboxyl or hydroxyl groups are able to take part in the ion exchange reaction in the following manner (Eqs. (2), (3) and (4))

n L–COOH + M n+ = n L–COOM + n H+(2)

n L–COOH + M n+ = (n? 1)L–COOM(OH)(3) Formation of hydroxilated species may be also considered

M n+ + H2O = M(OH)n + n H+(4) These reactions lead to the acidification of the solution after adsorption. The dissociation constant (p K a) of carboxylic groups are generally in between 4 and 6. Thus the increase of [H+] in solution is due to Eq. (1). Use of this material in removing the metal ions from acid mine wastewater has been investigated and the results are presented in this article. Present studies were aimed for those acid mine drainages which are contaminated by iron [Fe(II) and Fe(III)], manganese, calcium and not aluminum.

2. Materials and methods

2.1. Chemicals, materials and equipment

The lignite sample used in this study was obtained from Martin Lake, Texas. Prior to adsorption studies the samples were powdered to minus 325 B.S.S mesh. The chemicals were AR-grade. Stock solutions of the test reagents were made in de-ionized water from Millipipore-Q?. The pH was adjusted using dilute sulfuric acid and sodium hydroxide. The pH measurements were carried out using pH meter-model 710, Orion, USA. The metal concentrations in the samples were determined using Inductive Coupled Plasma (ICP) Spectrophotometer (Leeman PS 3000UV, USA). The carbon, hydrogen and nitrogen (CHN) analysis was carried out on LECO CHN analyzer, and XRD line profile analysis was performed using a Rigaku Geigerflex X-ray diffractometer.

3. Sorption procedure

3.1. pH optimization

The removal of Fe(II), Fe(III), Mn(II) and Ca(II) at different pH was studied in batch mode. Initial concentration of 100 ppm for Fe(II) and Mn(II) and 60 ppm for Fe(III) was used. A 25-ml of test solution of fixed concentrations was treated with 0.15 g of lignite and agitated intermittently for 48 h. The contact time and conditions were selected on the basis of preliminary experiments, which demonstrated that equilibrium was established in 48 h. After this period the solution was filtered using Whatman No. 42 and analyzed using ICP. The metal concentration retained in the sorbent phase (q e, mg/g) was calculated by using Eq. (5)

(5)

where C0 and C e are the initial and final (equilibrium) concentrations of the metal ion in solution (M), V the solution volume (l) and m is the mass of lignite (g).

3.2. Fixed bed studies

To determine the adsorption capacity of lignite under flow conditions, column experiments were performed in down flow mode. Both single and multi-columns were used to identify various design parameters.

The columns were setup according to the following procedure.

3.2.1. Single column

Approximately 10 g (dry) of (20 × 40 mesh) lignite was made into slurry with hot de-ionized water and fines were removed. The slurry was then fed slowly into the column such that the formation of air voids avoided/minimized. Glass wool was used as the column support. The fixed bed column was operated in the down flow mode. The adsorptive bed length was approximately 20-cm and the bed volume was 22-ml. A hydraulic flow rate of 2.5 ml/min was maintained using a peristaltic pump. Effluent samples were collected every 15 min, measured for pH and acidified using H2SO4(0.2 N) and stored for the chemical analysis. The columns were loaded with the metal ion solution containing the desired mixture of Fe(II), Mn(II), Zn(II), and Ca((II) [approximate concentration ratios = 60:50:50:50 ppm].

3.2.2. Multi-columns

The schematic representation of the setup is presented in Fig. 1. The columns were connected in series in down flow mode. Approximately 150 g (50-g in each column, dry) of lignite was weight (20 × 40 mesh) and made into slurry with hot de-ionized water and fines were removed. The procedure described above for filling single columns was repeated for all the columns used in this test. The metal solution was fed into the first column. The length of the adsorptive bed of each column was approximately 16 cm and the bed volume of the each column was 80 ml (total 240 ml).

A hydraulic flow rate of 10 ml/min was maintained using a Cole–Parmer Peristaltic Pump with master flex speed controller. Effluent samples were collected every 20 or 50 min from the third column, measured for pH and acidified using H2SO4 (0.2 N) and stored in 40 ml test tubes for the analysis.

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Fig. 1. Design of lignite fixed bed adsorbers (three bed system) used for the AMD treatment.

4. Results and discussions

Lignite is a member of the solid fuel family and is often referred to as brown coal. Macroscopically, lignite is dark brown or black in color when moist and light brown when dry. Its density lies in the range 1.0–1.35 g/cm3. Lignites are usually amorphous and fibrous or woody in texture. Lignite possesses high surface area, typically 100–200 m2/g [55] and [56].

Lignite was ashed to eliminate organic matter prior to determining the inorganic constituents. The ash was then acid digested and analyzed by plasma emission spectroscope. The chemical analysis of lignite shows the presence of SiO2 32.1%; Al2O3 13.4%; TiO2 1.30%; CaO 17.2%; MgO 1.81%; MnO 1.81%; Fe2O3; 18.8%; Na2O 0.28%; K2O 0.23%; BaO 0.14%; SrO 0.26%; C 55.56%; H 4.83%; N 1.14%; S 0.84% [2].

X-ray diffraction analysis of the sample indicated that the lignite is composed primarily of the amorphous phase with some quartz.

4.1. ZPC and leaching characteristics

Different amounts of lignite (2, 4, 8, 16, 32 mg/l) were added into de-ionized water having different initial pH (2–10) values. Blanks with no lignite were also run with the samples. All the experiments were performed in duplicate. The test tubes were agitated intermittently for 48 h at room temperature, and then left for settling of carbons. The final pH of the solution was measured. The variation in equilibrium pH at different lignite amounts is shown in Fig. 2. The equilibrium pH increases non-linearly in the pH range of 2–5. The increase is more pronounced in acidic medium

i.e. from 2.0 to 4.0. At higher initial pH there is no increase in the solution pH

[2]. The pH PZC was determined as the pH of the water that did not change after the contact with the samples. On the basis of this study, pH ZPC of lignite is considered to be pH 4.0. The equilibrium pH increases with increase in the amount of lignite at initial pH < pH PZC, whereas it decreases with increase in the amount of lignite at pH > pH PZC[2].

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Fig. 2. Effect of amount of adsorbent on the equilibrium pH of the de-ionized waters having no metal ions [2].

In view of the nature of the adsorbent used, the pH change in the system must be due to the exchange interaction of lignite in the solution

L2Ca + 2H+ = L2H2 + Ca2+(6)

L2Mg + 2H+ = L2H2 + Mg2+(7) These reactions are likely to occur from left to right in strongly acidic medium due to the high concentration of hydrogen ions. As result calcium, magnesium, sodium, aluminum and potassium

ions are released into the solution (Fig. 3). It is evident from Fig. 3that the leaching of these metals is directly proportional (except for Al3+, Fe2+) to the lignite's amount present in the water at varying pH values [2]. The change in concentration of Al3+, Fe2+was more complex due to simultaneous leaching and sorption of these ions [2].

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Fig. 3. Leaching behavior of lignite [2].

4.2. Effect of pH

The removal of Fe(II) and Mn(II) at different pH values is shown in Fig. 4(a)–(c). The sorption of Fe(II), Mn(II) and Fe(III) on lignite increases with the increase in pH. For sorption studies, the pH must be less than the pH for precipitation of respective metal ions. The sorption of Fe(II) is very low at pH in≤ 2, it increases from 6% to 84% (Fig. 4(a)) at pH 4.0. At pH in > 4.0 the removal takes place by sorption as well as precipitation i.e. the OH? ions from the solution formed some complexes with Fe(II). Also in the acid range pH fin increases with the increasing pH in, i.e. neutralization and sorption process are parallel processes. The optimum initial pH chosen for Fe(II) was 3.5, to correlate the removal with sorption process. The sorption of Mn(II) at different initial pH is shown in Fig. 4(b). Similar to Fe(II) the removal is negligible at low pH in≤ 2 and it increases with the increase in pH in and at pH in≥ 8.0 precipitation occurs [2]. Therefore the optimum pH chosen for sorption studies was 6.0. Also pH fin increases in more acidic pH in and after pH in4.0, the pH fin becomes more or less constant. Less removal of these ions at low pH have been attributed in this study to the competition between the protons and the metal ions. The removal of Fe(III) at different initial pH is shown in Fig. 4(c). The sorption is different from the other two metal ions. There is a sudden increase in the sorption process at pH 2.0, and pH in≥ 3.0 precipitation as well as sorption takes place [2]. The equilibrium pH also increases in the pH in range 2–3. The optimum initial pH chosen for the equilibrium studies was 2.7–2.8 to correlate removal with the adsorption process. Further, decrease in pH as a function of metal concentration is likely due to two equilibria:

Ion exchange reactions: n LH + M n+ = n LM + n H+(8)

Water hydrolysis: M n+ + n H2O = Mn(OH)n + n H+(9)