当前位置:首页 >> 环境科学/食品科学 >>

Bioremediation of Pb-contaminated soil by incubating with Phanerochaete 土壤重金属修复外文文献译文


通过加入黄孢原毛平革菌和稻草培养的方式对铅污染土壤的生物修复 实验室进行了利用培养好的白腐菌和秸秆对被铅污染的土壤进行生物修复模拟。监测了土壤的 pH 值,铅浓度,土 壤微生物,微生物代谢商,微生物商和微生物生物量 C 和 N 的比值。以上指标用来学习土壤中铅的强度和微生物 在生物修复过程中的影响。 研究表明被施以白腐菌和秸秆的土壤含有更低的可溶性交换铅, 更低的生物商和生物量 C 和 N 的比值(0 mg /kg 干土,1.9 mg CH2-C,生物量 C 和 4.9 在 60 天时) ,和更高的微生物生物量和微生物代谢 商(2258 mg /kg 干土和 7.86% 在第 60 天) 。另外,在 logistic 等式中的动力参数是用 BIOLOG 数据进行计算。对动 力参数进行分析后,就能得到一些微生物群的微生物量的信息。所有数据显示含铅土壤的生物利用度被减少,这样 潜在铅的强度被缓解,并且土壤微生物影响和微生物群的微生物量有所提高。 1. 简介

土壤中的重金属是最常见的环境污染。 铅被认定是所有重金属中危害最为严重的。 铅污染的主要来源是采矿、 冶炼、 含铅汽油、污水污泥、废弃电池以及其他含铅产品。这些种类繁多的铅来源导致土壤中含铅量偏高。Linet al 的报 道指出在瑞典 Falun 西南部大量工厂废物聚集地, 土壤含铅量超 1000 mg /kg。 Buatier et al.指出在法国一个污染地, 地表铅浓度达到 460–2670 mg/kg。铅的毒性和生物利用度受土壤 pH、氧化还原和铅种类的影响。土壤中的含铅化 合物主要通过可交换物、碳酸盐类、Fe/Mn 氧化物有机物和残留态流失。可溶性可交换状态铅的最大危害是铅非 常容易浸入地下水,地表以及农作物。然而铅在有机物和残留状态却无害,这是由于有机健的强度和硫化物,特别 是在重污染土壤中。因此,相对其他状态下的铅,铅在可溶状态时对环境,生态和人类更加有害。这样怎样减小土 壤中铅变为可溶状态是值得关注的。 相比传统的物理化学方法,生物修复是一种既不会加剧其他污染又能有效修复污染甚至还原土壤原先状态的技术。 之前的研究着重于物理方法研究含重金属土壤的微生物。 然而没有可用的信息用于向含稳定金属污染的土壤微生物 接种。但是重金属是非降解性污染并且很难在一些情况下用一般方法移除,因此总金属量很难大量减少。为了阻止 金属离子从土壤进入食物链或地下水, 要添加微生物吸附和积累金属离子作用于土壤中的固定污染金属。 之前的研 究表明白腐菌能够很好地吸收来自他稀释的金属和少量铅离子的转移。 白腐菌中能够积累在其细胞内的金属离子摄 取。正如大多数研究者所说,也可以与活跃细胞(包括死细胞)壁表面的官能团羧基,基或其他金属离子结合。同 时, 白腐菌能够在固体和液体环境中以及营养不足的环境中成长。 所以它能适应各种复杂的污染环境并且比其他微 生物成长更好。 此研究的目标是种白腐菌和秸秆在含铅土壤中去减少可溶的铅并提高生物活性。 系统分析了生物修复过程中的铅含 量和微生物指标的变化, 数据用来评估由孵化和无接种白腐菌中的铅污染土壤的修复效果。 这些结果预计对减轻金 属污染土壤接种白腐菌和秸秆对环境的影响提供有益的参考。 2. 材料和方法 2.1 微生物准备 采用白腐担子菌和含有 BKM-F-1767 的白腐菌。备用种在 4℃的条件下被保存在麦芽分解琼脂斜面上。在无菌蒸馏 水中制备孢子悬浮液。测量真菌浓度并将其调整至 2.0 × 106 CFU? ml?1。 2.2 土壤性质和孵育 在中国长沙岳麓山人迹罕至的山坡上,大量砾石和有机肥料被移除的地下 100cm 处收集未受污染的土壤。土壤是 自然风干并通过 2mm 尼龙网, 它的主要物理化学性质如下: 39%的粘土, 含有 0.83%有机碳, N 总量为 0.059%, PH 值为 4.9, 总的 Cu, Cd, Pb 分别为 11.5, 0 和 17.9 mg/ kg。 然后土壤和 Pb(NO3)2 溶液混合, 为了增加含 400mg Pb2+ 的孵育 5 周的干土,这样刺激含铅土壤成为相对稳定的状态。 2.3 实验设计 实验仪器包括试验用反应堆,二氧化碳移除器,加湿器,和降解产生的二氧化碳收集器。吹风机用于空气流动,空 气流动由流量计控制在 0.1 m3/h。空气流过 2M 的氢氧化钠时二氧化碳被移除。含纯净水的加湿器被用来阻止任何 碱性溶液进入反应器,并能够增加进入空气的湿度。反应器是 5L 的玻璃密闭瓶。被加湿无二氧化碳的空气从底部 的塑料孔进入反应器。氢氧化钠中的二氧化碳每三天更新一次。准备两组相同的实验仪器并标明 A 和 B。每组反应 器加入 1.5Kg 之前准备的土壤。每组反应器放入等量的秸秆,其和土壤的比值为 1:6,此混合物要被调整至 60%的含 水量。上述准备的孢子悬浮液要按 1:2 的重量比接种在 B 反应器中,A 不需要接种。A 中不含白腐菌污染的土壤被 添加的秸秆孵育, 而 B 中含白腐菌接种污染的土壤菌和秸秆被孵育。 剩余的秸秆提高了土壤孔隙度使其有更好的通 风并提供必要的代谢底物营养物质的微生物。多余的反应控制器中有土壤和白腐菌,并标为 C。在准备一个反应器

D.-L. Huang etal. / Journal of Hazardous Materials B134 (2006) 268-276 269 D,其中放入不接种也无秸秆的受污染的土壤。这样两组可以更好地看出内在固定土壤中铅的指标。两种土壤都要 培养 60 天。

2.4 土壤 pH 和铅的确定 土壤 pH 用摇晃 30 分钟 1:10 的水进行测量。铅的 5 个分数用一下表示: (i) 可溶性交换:1g 干土和 8ml 的 1 M MgCl2 (pH=7.0)分解 1 小时。 (ii) 碳酸类: (i)中的残留物和 pH 为 5 的 1 M 的 NaOAc 分解 5 小时。 (iii) Fe-Mn 氧化物: (ii)中的残留物和 0.04 M NH2OH?HCl 分解 6 小时。 (iv) 有机物: (iii)中残留物被加入 0.02M HNO3 和 30% H2O2,用 HNO3 将 pH 调为 2,混合液加热至 85°C 并保持两小时。再加入 3ml 过氧化氢。 (v) 残留态:减去其他四步所有的铅就是残留的铅。 2.5 微生物生理指标分析 这部分可以提供土壤中微生物化学的信息。和微生物 C(Cmic)的测量是用样品的熏蒸。土壤 qCO2 是土壤产生二氧 化碳和 Cmic 的比值。二氧化碳的产量是用测量的。样品置于 80 度烤箱中烘干,然后移至 550°C 排气管 5 小时。 碳键可以通过在点火时失去的重量被估测。微生物商是 Cmic 与 Corg 的比值。 2.6 BIOLOG 菌落生理分析 Kell 和 Tate 评估土壤微生物群落代谢多样性的潜在使用结构 BIOLOG。这表明碳的唯一土壤生物源利用率。5g 新鲜 的土壤增加 100 瓶的无菌水和动摇摇床为硝酸混合物加热至 85℃ 2 小时 10 分钟,10 倍系列稀释这种土壤。10-3 稀释(150μ L) ,然后到每一个结构 BIOLOG GN 板以及用于接种。这些板块在 25℃的 156h 培养。展色在 590 nm 处 测定光密度(OD) ,OD590 为每 12 小时的间隔读取。其他各井 OD590 减去 OD590 控制以及不含碳源。平均吸光度 (平均颜色的发展,AWCD) ,然后计算出每个板块,每个阅读时间,对时间 AWCD 曲线绘制。用动力学模型参数 和曲线拟合时间评估 AWCD 每个土壤样品。 2.7 数据分析 每次试验,分 3 组平行实验反应堆运行。取三个反应堆的平均值,用标准偏差来总结实验数据。BIOLOG 分析,SPSS 12.0 软件用于 Windows(SPSS 德国)的软件包,从获得的动力学参数进行统计分析。这些测试包括: (1)非线性 回归分析动力学参数值,并提出了密度依赖的 Logistic 生长曲线来描述每个土壤样品的 AWCD; (2)单向方差分析 (ANOVA)单一的时间点 OD 值和动力学模型参数。 3. 结果 3.1 土壤 PH 值随孵化的变化 在土壤培养的早期阶段,样品从反应器 A 和 B 的 pH 值略有下降。9 天之后,两个样品的 pH 值显著增加,然后趋 于稳定。pH 值上 60 天,A 变为 6.7,B 变为 7.3。在 B 土壤的 pH 值呈中性,高于在 A 的土壤。 3.2 土壤铅浓度和扩散系数(γ i) 所有土壤在孵化过程中的铅浓度变化。最高值在 A-? 土壤铅含量的可溶性交换显示的第 6 天,然后显著下降。经过 60 天的潜伏期,在 B 土壤铅可溶性交换浓度甚至下降到 0 mg/Kg,仍分别为 100.5 和 77.0mg/Kg,而在 C 土壤。结 果表明,其他四个 PB 的分数在 D 土壤对可溶性交换铅含量略有下降。经过 6 天的潜伏期,碳酸盐结合铅,有机结 合铅和 A 至 C 土壤中的残留铅明显增加,而铁锰氧化物限制的铅 18 天后增加。与对照土壤相比碳酸盐结合铅,有 机绑定铅和铅残留,和交换可溶性铅的最低浓度最高浓度,发现乙孵化后的土壤,它提供了 B 土壤中铅的最低的流 动性和铅的生物利用度。 3.3 微生物生理指标 微生物生物量,土壤有机质的生活的一部分,可以是一个很好土壤中铅的毒性的指标比较。在培养期间 Cmic 发生 了显著的变化。qCO2, CO2-C-Cmic 比例计算,高于乙土壤在整个孵化后 12 天。经过 6 天的潜伏期,Cmic 土壤有 机碳的比例, 远低于在 B 土壤。 在 B 土壤 Cmic/ Corg 第 24 天的最高值。 Cmic/ Corg 在 A 和 B 土壤的变化是相似的, 这表明对整个孵化期间的跌势。发现在土壤 B 中的 Cmic/ Nmic 整个孵化过程中的比例要低得多。 3.4 动力模型和参数

D.-L. Huang etal. / Journal of Hazardous Materials B134 (2006) 268-276 270 样品 AWCD 由 BIOLOG 和非线性与土壤样品接种的微孔板孵育时间决定,色彩的发展曲线的形状一般是 S 型,可以 通过描述基于密度依赖的 Logistic 生长方程的动力学模型。我们用特罗姆修改公式。

4. 讨论 以往的研究表明,重金属总量,不能反映金属的流动性和生物利用度,而有效浓度的金属与金属的毒性和工厂的可 用性有显着关系。同一种重金属,可溶性盐,在交换阶段最容易被植物吸收,因此可以通过观察在哪个阶段对存在 的金属进行评估。重金属的生物利用度和转移能力的降低与提取。经过 60 天的生物修复主要在 B 土壤铅的残留态 和有机较少的流动性和活动性, 而在土壤中的铅主要是铁锰氧化物和可溶性交换分数。 我们的研究结果表明生物体 的铅毒性,从铅合作孵化白腐菌中添加秸秆的土壤相比在 B 土壤铅的显著减少对环境的压力。 另一个可以解释为低铅后在 B 土壤活性生物修复的机制,是在 B 土壤较高的 pH 值。原土的 pH 值只有 4.9。pH 值 是影响离子的形式和化学的流动性。高 pH 值可能有助于减少金属在介质中的溶解度,这也是在我们的研究结果的 证实,少的水溶性可交换铅浓度的两种土壤中的 pH 值上升较低,土壤中活跃的铅浓度具有较高的 pH 值。可能是 由于日益增加的 pH 值有利于阳离子重金属保留土壤表面通过内部球体表面络合,吸附,沉淀和多核型反应。 guttormsen 等人发现,土壤 pH 值影响金属水解,离子对的形成,有机物的溶解度,以及表面电荷的铁和铝的氧化 物,有机质,粘土边缘。阿佩尔和马报道了 pH 值,土壤中重金属的吸附的重要作用,因为它直接控制金属氢氧化 物,以及金属碳酸盐和磷酸盐的溶解度,因此较高的 pH 值有利于降水和固定金属。所形成的氨溶液导致有机氮氨 化铵的形成和在土壤中的 pH 值的增加。 这也许可以解释在 A 和 B 的土壤 pH 值在孵化时间的观察。 在 B 土壤的 pH 值上升比在土壤 A 中快, 其原因可能是, 在 B 土壤菌体育促进有机物降解和氨或有机挥发溶酸。在 B 土壤有效的铅浓度低于对照组,这有利于减少毒性。 布鲁克斯建议,可以通过比较微生物参数评价重金属污染土壤生态系统功能的相对影响,越来越多的证据表明,土 壤菌群起着生态等级养分循环过程, 微生物的一个重要的角色比土壤生长在同一土壤重金属动物或植物都更为敏感。 因此,我们分析了一些微生物参数,以评估微生物的生长和微生物活动。微生物生物量是更为敏感的指标,比总有 机质含量变化的土壤条件和微生物生物合成的抑制。长期土壤菌群暴露于高重金属含量将减少 Cmic / Corg,因为金 属的毒性降低土壤微生物生物量和代谢效率。因此,减少铅胁迫对土壤菌群可以促进微生物的生长,可能是负责在 A 土壤和 B 土壤。Cmic 和 Cmic / Corg 增加,在孵化期间两种土壤的表示,在碳基板转换成生物代谢效率提高。 更高的 qCO2 在大部分金属污染土壤和 qCO2 大于未被污染的土壤中约 2.0 倍。傣族等也发现,qCO2 缓解金属毒性 并降低。12 天的生物修复,这可能表明,两种土壤中的铅毒性缓解后,在两种土壤 qCO2 显著下降。在 A 土壤中的 qCO2 比 B 土壤整治后大 1.6 倍。微生物代谢效率,抑制金属的存在,微生物合成需要更多的能量。主要是利用土 壤中有机碳,为维护能源金属污染的条件下生长的微生物,所以二氧化碳的释放增加,并转换成生物有机碳下降。 相反,微生物可以转换成生物基板,有效地减少或无金属污染。这些发现也许可以解释 qCO2 在我们的研究中观察 到的变化。 不同 C / N 比的微生物(如真菌和细菌)土壤生物量 C / N 比贡献。一般来说,有污染土壤和未被污染的土壤中的细 菌,真菌,因真菌耐受金属的主导地位。由于微生物 C / N 比值(如真菌,细菌等)之间的差异,在土壤中的生物 量,C / N 比变化与不同的微生物的不平等的增长。Cmic/ Nmic 显示减少 CMIC 增加,这可能是样品,因为低 C / N 比的微生物的增长速度比那些高 C / N 比的快。 约根森等人也证实, 是大量的微生物生物量 C / N 比值增加真菌重金 属污染下土壤微生物生物量比率造成的,他们报道的 C / N 为 3.5:1 和真菌,细菌的比率低,从 10 到 15:1。虽然 B 土壤比在孵化过程中的一个土壤 Cmic / Nmic 下降,我们的研究结果表明具有较高的污染水平和更高的的 Cmic / Nmic 土壤在 B 土壤 60 天的污染水平较低(低 Cmic 在土壤/ Nmic 表 1 和 2) ,这是以往的研究和我们上面得到的结 论相反。 Cmic/ Nmic 是不适合用来作为估计污染土壤中金属毒性和污染程度与微生物的接 Cmic 种,这可能是因为 Cmic / Nmic 接种的微生物会影响这个估计指数。 上面提到的微生物指标下铅污染的土壤中的微生物和微生物活动的增长反映的差异, 但不表明微生物群落状态的变 化。结构 BIOLOG 程序可以表明土壤微生物群落代谢能力。 微生物群落的代谢活动之间的差异可以通过分析对 Biolog 板碳源的利用进行评估。 整治后, 观察 B 土壤表明在我们 的研究通过共同培养法提高土壤中的微生物利用碳源的高 K 和低 S。有人还指出,有动力学参数 K 和 S 之间的显著 性差异,符合 Logistic 生长曲线在土壤和那些在 B 土壤,在 A 土壤中 AWCD 和显著性差异,而没有在 B 土被发现。 上述结果表明,动力学参数比 AWCD,是用来评估微生物群落的活动和土壤生态状况,按照以前的报告中的微生物 群落的代谢能力的变化更加敏感。 5. 结论 总之,我们的研究结果表明,孵化污染土壤接种白腐菌中,作为营养补充的秸秆在一起,可以减少活动的铅,减轻 铅应力,并稳定铅污染的土壤。此外,治疗比较对照组改善土壤的整治。所有这些结果可能是因为铅离子被吸收白

D.-L. Huang etal. / Journal of Hazardous Materials B134 (2006) 268-276 271 腐菌中的菌丝体,并在孵化过程中形成的腐殖质螯合。然而,进一步的研究需要进行调查和确认铅的固定机制的机 制。 依靠传统的生物修复技术对植物可能很难修复一些瘦肉与重金属污染的土壤, 因为土壤条件不利于植物的生长。 此外,当土壤中的微生物已经恶化,传统的物理化学方法不能有效改善土壤中微生物活动,但只从污染土壤中的金 属。在我们的研究中所使用的方法,不仅可以固定在土壤中的铅离子,而且还有效地提高土壤中微生物活动和微生 物群落的代谢能力。然而,只是固定金属离子络合,而不是从我们的研究土壤中移除,这样的方法需要改进且值得 进一步研究。

Bioremediation of Pb-contaminated soil by incubating with Phanerochaete chrysosporium and straw
Dan-Lian Huanga, Guang-Ming Zenga b'*, Xiao-Yun Jianga, Chong-Ling Fenga, Hong-Yan Yua, Guo-He Huang ab, Hong-Liang Liua
a

College of Environmental Science and Engineering, Hunan University, Changsha 410082, Hunan, China b Faculty of Engineering, University ofRegina, Regina, Canada S4S 0A2 Received 22 July 2005; received in revised form 5 November 2005; accepted 8 November 2005 Available online 15 December 2005

Abstract
The bioremediation of the simulated lead (Pb)-contaminated soils by incubating with Phanerochaete chrysosporium and straw was studied at laboratory-scale. The soil pH, Pb concentration, soil microbial biomass, microbial metabolic quotient, microbial quotient and microbial biomass C-to-N ratios were monitored. The above indicators were to study the stress of Pb on soil and the microbial effects during the bioremediation process. It was found that the soils treated with P chrysosporium and straw showed a much lower concentration of soluble-exchangeable Pb, lower metabolic quotient and biomass C-to-N ratios (0 mg kg-1 dry weight soil, 1.9 mg CO2-C mg-1 biomass carbon and 4.9 on day 60, respectively) and higher microbial biomass and microbial quotient (2258 mg kg-1 dry weight soil and 7.86% on day 60, respectively) compared with the controls. In addition, the kinetic parameters in the model based on logistic equation were calculated by the BIOLOG data. By analyzing those kinetic parameters some information on the metabolic capacity of the microbial community could be obtained. All the results indicated that the bioavailability of Pb in contaminated soil was reduced so that the potential stress of Pb was alleviated, and also showed that the soil microbial effects and the metabolic capacity of microbial community were improved. ? 2005 Elsevier B.V. All rights reserved.
Keywords: P chrysosporium; Inoculation; Bioremediation; Pb-contaminated soil

* Corresponding author. Tel.: +86 731 8822754; fax: +86 731 8823701. E-mail address: zgming@hnu.cn (G.-M. Zeng).

1. Introduction
Heavy metal contamination in soil becomes a widespread problem. Lead (Pb) has been recognized as one of the most hazardous heavy metal among environmental pollutants. The primary sources of Pb-contamination come from the mining and smelting activities, combustion of D.-L. Huang etal. / Journal of Hazardous Materials B134 (2006) 268-276 269 leaded gasoline, land application of sewage sludge, battery disposal and Pb-bearing products [1]. Such irregular inputs of Pb result in the high -1 concentrations of Pb in soils. Lin et al. [2] reported that soils with Pb concentration higher than 1000 mg kg occurred in the south-western part of Falun, Sweden, where large amount of the industrial wastes was deposited. Buatier et al. [3] found that the Pb concentrations were -1 460-2670 mg kg in the surface horizons of a polluted site in France. The toxicity and bioavailability of Pb are affected by soil pH, redox potential and Pb species. It is commonly accepted that Pb compounds in soil mainly exist in exchangeable, carbonate-bound, Fe/Mn oxide-bound, organic and residual phases [4,5]. The serious negative impacts of Pb in the soluble and exchangeable phases easily leached out of soils on groundwater and surface water and even crops have been observed, while Pb in the organic and residual phases is inactive due to the strong binding capacity of organic matter and sulfides, especially in the heavily contaminated soils [5,6]. Consequently, active Pb in soluble and exchangeable phases poses more threat to environment, ecosystem and human, compared with immobilized Pb in other phases, and how to reduce active Pb effectively receives much more concerns in the remediation of Pb-contaminated soils [6]. Comparing with conventional physico-chemical approach, bioremediation is a technology not to aggravate other environmental problems but to remediate the polluted soil partially or fully to the original state, which relies on the natural soil community or addition of exogenous organism (plant or microorganism) with exceptional metal-binding capacity to remove the heavy metals from soil or alleviate the toxicity of metals by reducing the bioavailability and mobility of metals largely. Previous studies focused on the bioremediation of soils contaminated with heavy metals (such as Pb, Cd, Cr and so on) by phytoremediation [7,8], however, little information is available on the application of inoculum of microorganism to stabilize metal-contaminated soils. But in fact, heavy metals are non-biodegradable pollutants and not easily removed from soils by regular treatments under some conditions, so the total metal content is difficult to decrease greatly. To prevent metal ion in soil from entering food chain or groundwater, microorganisms adsorbing and accumulating metal ion are expected to be applied to immobilize metals in soils. In previous studies, it was proved that P. chrysospo- rium was good at absorbing metal from dilute solutions by its mycelium and less Pb ion transferred [9]. P chrysosporium are capable of accumulating metal ions in their cells by intracellular uptake, as many researchers validated, and can also be chelated with metal ions by the carboxyl, hydroxyl or other active functional groups on cell (including the dead cell) wall surface [10]. Meanwhile P chrysosporium is able to grow in both solid and liquid environment and degrade a wide range of xenobiotic effectively even in the nutrient-limited condition [11]. So it could be adapted to complex polluted environment and grow better than other microorganisms after inoculation into soils. As a result, it is of its own advantage to be applied to the bioremediation of metal-contaminated soil. But there are few reports about its application to inactivate metals in soils. The aim of this study was to inoculate P. chrysosporium as the exogenous microorganism, together with some straws, into Pbcontaminated soil for reducing the solubility and bioavailability of Pb and improving the soil microbial activity. The changes of Pb content and microbial indices that took place in bioremedia- tion process were analyzed systematically, all of which were used to evaluate the remediation effects on Pb-contaminated soils by incubating with and without inoculum of P. chrysospo- rium. These results are expected to provide useful references on alleviating the environmental impact of metal-contaminated soils by inoculation with P. chrysosporium and straw.

2. Materials and methods
2.1. Microorganisms preparation The white-rot basidiomycete, P chrysosporium strain BKM- F-1767 was used. Stock cultures were maintained on malt extract agar slants at 6 -1 4 °C. Spore suspensions were prepared in sterile distilled water. The fungal concentration was measured and adjusted to 2.0 x 10 CFUml . 2.2. Soil properties and preincubation Uncontaminated soils were collected from about 100 cm underground on the unfrequented hillside of Yuelu Mountain (Changsha, China), from which gravels and large organic scraps were removed. The soil was air-dried and ground to pass through a 2 mm nylon screen, and its main physico-chemical character-

Diyness NaOH Wate r Reacto r

Fig. 1. Schematic diagram of experimental apparatus.

istics were measured and listed as follows (dry weight): 39% of clay, an organic C content of 0.83%, total N of 0.059%, a pH value of 4.9, and -1 the total contents of Cu, Cd, Pb were 11.5, 0 and 17.9 mg kg . Then this soil was mixed and homogenized with Pb(NO 3 ) solution for adding 2 Pb + 400 mg kg dry weight soil, which was preincubated for 5 weeks so that the stimulated Pb-contaminated soil became relatively stabile.
2 1

2.3. Experimental design
0304-3894/$ - see front matter ? 2005 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2005.11.021

The experimental apparatus used for this research consisted of a lab-scale column reactor, a CO2 removal trap, a humidifier, and a trap for 3 -1 collecting CO2 generated in biodegradation as shown in Fig. 1. Blower was used for aeration, and the air flow was controlled at 0.1 m h by

D.-L. Huang etal. / Journal of Hazardous Materials B134 (2006) 268-276

270

flow meter. CO2 was removed from the incoming air by reacting with an alkali solution of 2 M sodium hydroxide (NaOH) so that the collected CO2 was entirely from decomposition [12]. The humidifier containing distilled water was used to prevent any aspirated alkali solution from entering the compost reactor, and to raise the moisture content of the incoming air. Air-tight glass vessel of 5 l was used as reactor. The air with CO2 removed and being humidified entered the reactor from bottom through perforated plastic plate. Emitted CO 2 was continuously trapped in a solution of 2 M NaOH which was renewed every 3 days. Two identical sets of experimental apparatuses were prepared and labeled as Reactors A and B. Each reactor was set up by adding 1.5 kg of the contaminated soil as prepared above. Equal straws were put into each reactor to mix thoroughly with soil in the ratios of 1:6, and this mixture was adjusted to 60% water content. The spore suspension prepared as above was inoculated in the weight ratio of 2% into the mixture in Reactor B, and Reactor A without inoculum was performed as control. Contaminated soil without P. chrysosporium was co-incubated with the added straws in Reactor A, while the contaminated soil with inoculum of P. chrysosporium was co-incubated with the added straws in Reactor B. The addition of straws could enhance soil porosity for better aeration and offer metabolizable substrate as necessary nutriments for microorganisms. The additional control reactor with soil and P. chrysosporium only was prepared and labeled as Reactor C. The Reactor D was also prepared with the contaminated soil, but without inoculum and straw, and is thus, another control. Such two controls were used for better indication of the intrinsic immobilization of Pb in soil. Both soils were incubated 60 days in this study. Five individual 10 g fresh samples were taken from different sites in the reactor periodically and mixed together homogeneously for routine analysis. Some of the samples were air-dried, loosed, and passed through a 2 mm nylon sieve. After removal of the straw the samples were used for Pb content measurement. On days 0 and 60, additional samples (three 10 g samples) were taken from reactor and subjected to homogenization for BIOLOG analysis. All the analyses were performed in duplicate. 2.4. Soil pH and Pb content determinations Soil pH value was measured in 1:10 sample-H2O extract after shaking for 30 min [13]. Because the toxicity of metal was associated with metal bioavailability, the soils were analyzed for Pb-fractions according to the continual extract procedure of Tessier et al. [14]. Five fractions of Pb were extracted in turn as follows: (i) Soluble-exchangeable: The 1 g dry soil was extracted for 1 h with 8 ml of 1M MgCl2 (pH 7.0) with continuous agitation. (ii) Bound to carbonates: The residue from (i) was extracted for 5 h with 8 of 1M NaOAc adjusted to pH 5.0 with acetic acid (HOAc). ° (iii) Bound to Fe-Mn oxides: The residue from (ii) was extracted for 6 h with 20 of 0.04 M NH 2OHHCl in 25% (v/v) HOAc at 96 C with occasional agitation. (iv) Bound to organic matter: The residue from (iii) was added 3 of 0.02 M HNO3 and 5 of 30% H2O2 adjusted pH 2 with HNO3, and the mixture was heated to 85 °C for 2h with occasional agitation. A second 3 ml aliquot of H 2O2 (pH 2 with HNO3) was then added and the sample was heated again to 85 °C for 3 h with intermittent agitation. After cooling, 5 of 3.2M NH4OAc in 20% (v/v) HNO3 was added, and the sample was diluted to 20 and agitated continuously for 30 min. (v) Residual: The residual Pb was calculated by subtracting the other four fractions from total Pb. Additional 1 g dry soil was digested with a 5:1 mixture of hydrofluoric and perchloric acids for total Pb measurement. Between each successive extraction, separation was effected by centrifuging and Pb concentrations in the supernatant were determined by an atomic absorption spectrometer (Agilent 3510, USA). The distribution coefficient Y I of each Pb-fraction showed the ratios of Pb content in the different fractions and was calculated as follows:

In Eq. (1), fi and f represent content of each Pb-fraction and total amount of Pb in soils, respectively. 2.5. Microbial physiological indices analysis Studies on soil microbial biomass carbon (Cmic), microbial metabolic quotient (qCO2), microbial quotient (Cmic/Corg) and microbial C-to-N ratio (Cmic/Nmic) can provide information on the biochemical processes occurring in the soil, and there is growing evidence that soil biological parameters may be of great potential as the early and sensitive indicators of soil ecological stress and restoration [15-17]. So we analyzed these microbial physiological indices listed above. The Cmic and biomass N (Nmic) were measured by fumigation of the sample with ethanol-free chloroform and extraction with 0.5M K2SO4, according to Brooks et al. [18] and Vance et al. [19]. The soil qCO 2 was the ratio of CO2 produced by soil basal respiration and Cmic [20]. The CO2 production was measured by titration with 1M HCl after adding an excess of barium chloride and phenolphthalein indicator to the alkali solution trapping CO2 as previously described. The samples were dried to constant weight in the oven at 80 °C, and then ignited in muffle at 550 °C for 5h. Afterwards, the organic carbon (Corg) could be estimated from the organic matter content calculated by the weight loss on ignition, as described by Huang et al. [21]. The microbial quotient was calculated by the ratio of Cmic to Corg (Cmic/Corg). 2.6. BIOLOG community-level physiological profiling

D.-L. Huang etal. / Journal of Hazardous Materials B134 (2006) 268-276

271

The potential metabolic diversity of soil microbial communities was assessed using BIOLOG plates as described by Kell and Tate [22], which showed the sole carbon source utilization of soil biota. Fresh soil equivalent to 5 g dry weight was added to 100 sterile water in a flask and -3 shaken with a rotary shaker for 10 min, and 10-fold serial dilution was made for this soil suspension. The 10 dilution (150 ^l) was then used to inoculate into each well of BIOLOG GN plate. These plates were incubated at 25 °C for 156 h. Color development was measured as optical density (OD) at 590 nm, and the OD590 was read for each well at 12 h intervals. The OD 590 of control well containing no carbon source was subtracted from the OD590 of each of the other wells. The average absorbance (average well color development, AWCD) was then calculated for each plate at each reading time [23] and plotted against time to give AWCD curves. To evaluate AWCD of each soil sample, kinetic model parameters and curves fitting to the color development time-course data were also considered. 2.7. Statistical analysis For each treatment, three parallel sets of experimental reactors were run. The results to be presented were mean value for the three reactors, and the standard deviations were used to summarize experimental data. Statistical analyses were performed on the kinetic parameters obtained from BIOLOG analysis, which used the software package SPSS 12.0 for Windows (SPSS, Germany). These tests included: (1) non-linear regression analysis performed to describe the AWCD of each soil sample for showing the kinetic parameter value and suggesting a density-dependent logistic growth curve and (2) a one-way analysis of variance (ANOVA) for the single time-point OD value and kinetic model parameters.
Fig. 2. pH changes in A and B soils during incubation: (A) control without inoculum and (B) treatment with inoculum and straw. The bars represent the standard deviation of the means (n = 3).

3. Results
3.1. Soil pH changing with incubation In the early stages of soil incubation, pH of samples from Reactors A and B decreased slightly. After 9 days, the pH values of both samples increased significantly and later tended to stabilize (Fig. 2). On day 60, pH turned to be 6.7, 7.3 for the Reactors A and B, respectively. The pH value in B soil was neutral and higher than that in A soil (Fig. 2). 3.2. Soil Pb concentrations and distribution coefficient (Yi) Pb concentrations in all soils varied during incubation (Table 1). The soluble-exchangeable Pb content in A-C soils displayed the highest value on day 6, and then remarkably decreased with incubation. After 60 days of incubation, the concentration of soluble-exchangeable Pb in B soil -1 -1 even dropped to 0 mg kg , while that in A soil and C soil remained 100.5 and 77.0 mg kg , respectively. It was shown that in D soil a slight decrease on the content of soluble-exchangeable Pb and a few changes in the other four Pb-fractions. The carbonate-bound Pb, organic- bound Pb and residual Pb in A-C soils increased obviously after 6-day incubation, whereas Fe-Mn oxides-bound Pb were found to increase after 18 days. Compared with the control soils, the highest concentration of carbonate-bound Pb, organic-bound Pb and residual Pb, and the lowest concentration of soluble- exchangeable Pb were found in B soil after incubation (Table 1), which offered the evidence for the lowest mobility and bioavail- ability of Pb in B soil. The higher the distribution coefficient (Yi) of each Pb- fraction, the more the corresponding Pb-fraction will be. The Yi of soluble-exchangeable Pb on day 0 was the highest among all Yi of Pb-fraction on day 0 in both A and B soils, which showed Pb existed mainly in the soluble-exchangeable form in both soils before remediation (Fig. 3). After 60-day incubation, in both A and B soils, the Yi of soluble-exchangeable Pb reduced, whereas that of the other four forms increased, which showed one form of Pb had been transformed into another (Fig. 3). It can be also observed that the order of abundance of the five fractions of Pb in B soil on day 60 was residual Pb >

D.-L. Huang etal. / Journal of Hazardous Materials B134 (2006) 268-276

272

organic-bound Pb > Fe-Mn oxides-bound Pb > carbonate-bound Pb > soluble-exchangeable Pb. These results revealed that the active Pb in B soil had been transformed into inactive Pb indicating the alleviation of Pb- contamination after 60-day remediation with P chrysosporium and straw. 3.3. Microbial physiological indices Microbial biomass, being the living part of soil organic matter, can be a good index for comparing the toxicity of Pb in soil. The C mic changed significantly during the incubation period (Table 2). It was also observed that the C i was much higher in B soil relative to that in A soil throughout after 6 days of incubation (Table 2). The qCO 2, as calculated by the CO2-C- to-Cmic ratio, was higher in A soil compared to that in B soil throughout after 12-day incubation (Table 2). After 6-day incubation, the C ic/C g, being the ratio of the C i to soil organic carbon, was much lower in A soil than that in B soil. And the C ic/C g in B soil displayed the highest value on day 24 (Table 2). The change of C ic/Nmic in A and B soils was similar, which showed a downtrend on the whole during incubation. Much lower C ic/N i ratio throughout incubation process was found in A soil (Table 2), compared with that in B soil.
mc m or mc m or m m mc

3.4. Kinetic model and parameters The samples' AWCD determined by BIOLOG were nonlinear with the incubation time for microplates inoculated with the soil samples, and the shape of the color development curve was generally sigmoidal (Fig. 4), which can be described by the kinetic model based on the density-dependent logistic growth equation. We used the equation modified by Lindstrom et al. [24]
Soil sample Fig. 3. Distribution coefficient (Yi) of Pb in A soil on day 0, B soil on day 0, A soil on day 60 and B soil on day 60, respectively: (A) control without inoculum and (B) treatment with inoculum and straw.

D.-L. Huang etal. / Journal of Hazardous Materials B134 (2006) 268-276 Table 1 Pb content of five fractions in A-D soils during incubation Pb-fractions Time (days) 0 6
1 a

273

12 238.4 (2.1) 200.5 (2.0) 236.6 (2.5) 252.4 (3.1)

18 213.1 (1.2) 154.3 (1.1) 192.1 (2.6) 240.9 (2.3)

24 183.6(1.7) 96.2 (1.7) 155.3 (1.9) 231.3 (2.2)

30 162.9 (4.2) 54.4 (1.3) 131.3 (2.0) 220.2 (3.4)

36 137.9 (3.8) 29.3 (1.0) 112.4 (2.7) 211.5 (3.6)

42 121.1 (0.5) 12.5 (0.7) 93.5 (1.2) 203.5 (2.8)

50 104.5 (0.8) 0.0 (0.2) 81.2(0.9) 201.7 (2.3)

60 100.5 (1.3) 0.0 (0.1) 77.0 (1.7) 200.1 (2.2)

Soluble-exchangeable Pb (mg kg ) A 254.7 (1.5) 259.0(1.9) B 258.9 (1.8) C 260.1 (2.6) D 259.4 (2.1) Carbonate-bound Pb (mg kgA 16.7 (0.4) B 12.5 (0.3) C 18.0 (0.6) D 17.2 (0.4) 266.7(1.1) 261.3 (2.4) 260.2(1.3)
1 a

)

8.4 (0.2) 4.2(0.1) 12.1 (0.3) 15.7 (0.5)

12.6 (0.3) 37.6 (0.4) 17.4 (0.6) 17.0 (0.4)

29.3 62.6 31.4 18.2

(0.1) (0.6) (0.8) (0.5)

33.4 (0.3) 87.8 (0.9) 34.0 (0.7) 19.9 (0.6)

45.9 (0.2) 66.9 (0.3) 32.2 (0.4) 18.4 (0.2)

50.2 (0.2) 71.1 (1.2) 33.2 (0.3) 19.1 (0.1)

41.8 (0.1) 66.9 (1.1) 34.9 (0.5) 20.7 (0.2)

29.3 (0.1) 54.2 (0.5) 28.7 (0.6) 19.6 (0.4)

25.1 45.9 24.1 19.1

(0.2) (0.8) (0.5) (0.4)

Fe-Mn oxides-bound Pb (mg kg-1 )a A 79.3 (1.4) 87.7(1.6) B 79.3 (0.7) 100.0(1.8) C 80.3 (2.2) 88.0(2.1) D 80.9 (2.0) 82.2(1.7) Organic-bound Pb (mg kg-1 )a A 54.3 (0.2) B 58.5 (0.5) C 53.7 (0.4) D 52.0 (0.3) 46.0 (0.6) 37.5 (0.1) 51.2(0.2) 50.9 (0.4)

75.3 (0.9) 71.0(1.2) 71.8 (1.5) 80.6(1.4) 46.0 (0.4) 45.9 (0.3) 57.1 (0.6) 50.7 (0.3)

66.9 45.9 64.1 81.0

(1.2) (0.4) (1.3) (1.8)

70.9 (1.3) 62.7 (0.8) 76.0 (1.6) 79.9 (1.3) 58.4 (1.0) 66.9 (1.1) 78.7 (1.3) 55.7 (1.1) 70.9 (1.3) 104.5 (1.1) 77.2(1.7) 32.9 (0.6)

83.5 (1.9) 87.8 (1.4) 86.4 (1.5) 83.2(1.9) 62.6(1.4) 83.7 (1.8) 84.8 (2.0) 57.8 (1.2) 62.6(1.7) 125.5 (1.8) 86.8 (1.4) 40.0(0.5)

100.3 (1.6) 87.9 (0.6) 87.4 (2.0) 83.5 (1.2) 71.0(1.9) 104.6 (2.0) 100.5 (2.2) 58.2 (0.9) 58.5 (1.2) 125.5 (2.1) 87.5 (1.6) 47.4 (0.7)

116.9 (2.1) 96.1 (1.3) 92.4 (1.9) 85.0(1.5) 75.2(1.5) 112.8 (1.9) 106.4 (2.1) 60.8 (1.4) 62.6(1.4) 129.6(2.2) 93.3 (1.1) 49.6(0.9)

129.5 (2.3) 104.2 (2.5) 96.4 (1.4) 86.2 (1.0) 83.6 (1.6) 120.9 (2.3) 113.7 (2.5) 61.9 (1.2) 71.0(1.1) 137.5 (2.4) 100.8 (1.2) 50.4 (0.8)

134.0 108.5 100.3 86.9

(1.0) (1.2) (1.5) (1.1)

50.1 (0.9) 54.2 (0.7) 65.9 (1.1) 52.4 (0.8)

87.9 (0.8) 121.0 (1.3) 116.3 (1.2) 62.3 (0.8) 71.2(0.9) 141.9 (1.7) 103.2(1.0) 51.2(0.6)

Residual Pb (mg kg 1 )a A 12.5 (0.1) 16.7(0.2) 46.0(0.5) 58.5 (1.6) B 8.4 (0.1) 8.3 (0.2) 62.6(0.9) 100.1 (1.3) C 9.1 (0.3) 8.9(0.1) 39.2(0.6) 68.1 (1.2) D 10.1 (0.2) 10.8(0.4) 19.0(0.5) 27.2(0.9)

(A) Control without inoculum; (B) treatment with inoculum and straw; (C) control without straw; (D) control without inoculum and straw. a Values are means (n = 3) with standard deviation in parentheses.

as below, which is a form providing parameters being interpreted more readily with respect to the shape of the OD kinetic curve and to:
D.-L. Huang etal. / Journal of Hazardous Materials B134 (2006) 268-276 274

y = OD590nm =

K 1+e underlying microbiological behavior driving its shape. In Eq. (2), K represents the asymptote (y = K) approached by the mean test well OD curve that means the highest OD in the culture course, the unitless coefficient r determines the exponential rate of OD change, and t is the time following inoculation of the microplate, the exponential parameter s in the denom
-r(t-s)

Table 2 The microbial parameters of A and B soils during incubation Time (days) Biomass carbon (mgkg 1 soil)a Metabolic quotient (mg CO2-Cmg 1 Cmic)a Biomass C/organic C (%) a Biomass C/Na

A 0 3 6 9 12 15 18 24 27 30 36 42 50 60 979(10) 840 (8) 892(12) 965(15) 1158(31) 1397 (54) 1547 (55) 1364 (67) 1276(62) 1331 (59) 1301 (48) 1379 (75) 1351(64) 1311(73)

B 1034(34) 763 (17) 991 (26) 1535 (79) 1714(51) 1935 (76) 2473 (93) 2789(68) 2271 (72) 2171(80) 2090(55) 2297 (83) 2287(69) 2258(81)

A 4.2 (0.5) 4.4 (0.4) 4.5 (0.8) 5.4 (1.1) 4.6 (1.0) 3.7 (0.7) 3.7 (0.3) 4.0 (0.6) 3.7 (0.4) 3.5 (0.5) 3.2 (0.3) 3.1 (0.4) 3.1 (0.2)

B 4.5 (0.4) 5.1 (0.9) 5.5 (0.7) 3.9 (0.2) 3.3 (0.2) 2.7 (0.4) 2.1 (0.1) 2.5 (0.3) 2.6 (0.5) 2.2(0.1) 2.0 (0.2) 1.9 (0.3) 1.9 (0.1)

A 2.12(0.18) 1.89 (0.22) 2.23 (0.29) 2.55 (0.35) 2.98 (0.38) 3.65 (0.42) 3.91 (0.31) 3.43 (0.42) 3.34 (0.26) 3.78 (0.23) 3.69 (0.36) 4.04 (0.40) 4.31 (0.19) 4.50 (0.25)

B 2.27 (0.24) 1.71 (0.13) 2.44(0.18) 3.99 (0.49) 4.57 (0.33) 4.96 (0.48) 6.20 (0.54) 8.46 (0.73) 7.25 (0.70) 6.60 (0.55) 6.10(0.44) 6.82 (0.67) 7.47 (0.36) 7.86 (0.43)

A 7.5(1.1) 7.9 (0.9) 8.3 (1.2) 8.9 (0.7) 8.1 (0.6) 6.8 (0.5) 5.9 (0.8) 6.4 (0.7) 6.6 (0.9) 5.7 (0.3) 6.1 (0.4) 4.9 (0.2) 4.5 (0.1) 4.2 (0.2)

B 8.9 (1.2) 9.3 (1.5) 10.2 (0.5) 10.3 (0.4) 9.5 (0.3) 8.7 (0.7) 8.2 (0.6) 7.4 (0.8) 7.7 (0.2) 6.6 (0.4) 6.8 (0.5) 5.6 (0.3) 5.1 (0.3) 4.9 (0.1)

nator is the time to the midpoint of the highest mean OD Each plot in Fig. 4 presented the mean OD change over
A soil on day 0

(2)

(when y = K/2). time in the blank-corrected test wells of

the three replicate

0 24 36 48 60 72 84 96 108 120 132 144 156 Incubation time (h)

D.-L. Huang etal. / Journal of Hazardous Materials B134 (2006) 268-276

275

60 72 84 96 108 120 132 144 156 Incubation time (h)

A soil on day 60

0 24 36 48 60 72 84 96 108 120 132 144 156 Incubation time (h)

B soil on day 60

K = 1,49

r=0.09l
=50.9 RJ=0.998 7

j

0 24 36 48 60 72 84 96 108 120 132 144 156 Incubation time (h)

plates. The curve fitting technique fitted the time-course OD data for microplate test wells closely and was useful for estimating kinetic parameter data that reflect the response of culturable organisms in the microplate 2 inoculum. All kinetic parameter values from fit curves fitting to the OD time-course provided a good (R > 0.99) fit to the sigmoidal kinetics of color development data (Fig. 4). The parameters K and r values estimated for A soil and B soil on day 60 were much higher than those on day 0, and the parameter K for B soil increased more than that for A soil. The parameter s for A and B soils was reduced from days 60 to 29.4 and 30, respectively, compared with that on day 0 (Fig. 4), which meant much less time needed to reach the midpoint of the highest mean OD in B soil after bioremediation. These results showed the improvement of the metabolic activity of microbial community in A and B soils by remediation, and the microbial community in B soil was of better metabolic capacity after bioremediation with P. chrysosporium and straw, compared with that in A soil as control (Fig. 4). One-way ANOVA was performed to the OD time-course data and the kinetic parameters generated by the model. It is found that there was significant difference (P <0.05) between the kinetic model parameters K and s in A soil on day 60 and those in B soil, whereas the OD time-course data and the r in A soil on day 60 was not significantly different from (P = 0.128 and 0.293, respectively) those in B soil.

4. Discussion
Previous studies have shown that the total amount of heavy metals cannot reflect the mobility and bioavailability of metals well, whereas the effective concentration of metal has significant relationships with the

toxicity and plant availability of metal [25,26]. To the same kind of heavy metal, soluble salt in exchangeable phase is the easiest to be assimilated by plants and as a result the toxicity of Pb in soil to environment can be evaluated by observing in which phase the metal existed [27]. The bioavailability and transfer ability of heavy metal is reduced with the turn of extraction [14]. After 60-day bioremediation the Pb in B soil was mainly bound D.-L. Huang etal. / Journal of Hazardous Materials B134 (2006) 268-276 to residual fraction and organic fraction with less mobility and activity, while the Pb in A soil was mainly bound to Fe-Mn oxides fraction and soluble-exchange fraction. Our results showed the least toxicity of Pb to living organisms, the least stress from Pb on environment for the significant reduction of active Pb in B soil by co-incubating the soil with P. chrysosporium and the added straws, compared with those in the control soils (Fig. 3 and Table 1). Reasons for these results could be as follows: (i) white-rot fungi inoculated into B soil can be chelated with Pb by the carboxyl, hydroxyl or other active functional groups on cell wall surface to reduce Pb activity [9] and (ii) white-rot fungi could improve the organic

276

Fig. 4. Kinetics of average well colordevelopmentforAand B soils: (A)control without inoculum and (B) treatment with inoculum and straw on days 0 and 60. Absorbance values (—?—) are the measured values. Bars are standard deviation of the means (n = 3). The solid line (—) is a plot of the equation fit to the mean absorbance data. Kinetic model K, r and s parameter data for the fit equation are presented along with the R2 value of the fit to the absorbance data for each curve.

matter decomposition and nutrient cycling, as reported previously [21], which promoted the formation of humus, while that Pb was chelated with humus is the mechanism responsible for Pb immobilization. Another mechanism that could account for the lower Pb activity in B soil after bioremediation was the higher pH in B soil (Fig. 2). The pH of the original soil (no addition) was only 4.9. The pH is known to affect the ionic form and chemical mobility. A high pH may facilitate the decrease of the solubility of metals in the medium [28], which was also validated by the findings in our research that the less soluble-exchangeable Pb concentration was found as the pH in both soils increased (Table 1 and Fig. 2). The lower active Pb concentrations in soil with higher pH might be due to the increasing pH, which facilitates the cationic heavy metal retention to soil surfaces via adsorption, inner sphere surface complexation, and/or precipitation and multinuclear type reactions [29]. Guttormsen et al. [30] found that soil pH affected metal hydrolysis, ion-pair formation, organic matter solubility, as well as surface charge of iron and aluminum oxides, organic matter, and clay edges. Appel and Ma [31] reported that soil pH plays a major role in the sorption of heavy metals as it directly controls the solubilities of metal hydroxides, as well as metal carbonates and phosphates, and as a result the higher pH facilitates the precipitation and immobilization of metals. The solubilization of the ammonia formed by organic nitrogen ammonification led to the formation of ammonium and an increase in the pH values in soil. This might explain the increasing pH observed in both A and B soils during incubation time (Fig. 2). The pH in B soil increased more than that in A soil (Fig. 2), the reason for which might be that P. chrysosporium in B soil facilitated the degradation of organic matter [21] and the solubilization of ammonia or the volatilization of organic acid. The effective concentration of Pb in B soil was lower than the controls (Table 1), which benefited the reduction of toxicity. Brookes [32] suggested that the relative effects of heavy- metal contamination of soils on ecosystem function can be evaluated by comparing microbial parameters, and an increasing body of evidence suggests that soil microflora plays an important role in ecosystemlevel nutrient cycling processes and microorganisms are far more sensitive to heavy metal stress than soil animals or plants growing on the same soils [33-35]. So we analyzed some microbial parameters to evaluate the growth of microorganisms and the microbial activity in our study. Micro- bial biomass is a much more sensitive indicator of changing soil conditions than is the total organic matter content, and micro- bial biomass synthesis is inhibited in heavy metal-contaminated soils generally [36]. Long-term exposure of soil microflora to high heavy metal concentrations would decrease the C mic/Corg, because the toxicity of metal reduced microbial biomass and metabolic efficiency [37,38]. So the reduction of Pb stress on soil microflora could facilitate the growth of microorganisms and might be responsible for the increase of Cmic and Cmic/Corg in A soil and B soil (Tables 1 and 2). The increasing C mic and Cmic/Corg in both soils during incubation (Table 2) indicated that the metabolic efficiency in the conversion of carbon substrates into biomass increased. Our data also showed the Cmic and the Cmic/Corg were higher obviously in B soil than those in A soil (Table 2), which suggested the higher metabolic efficiency in carbon mineralization in B soil. With the addition of P chrysosporium, the total amount of soil microorganisms in B soil increased (Table 2). The inoculated P. chrysosporium could accumulate metal on cell wall and deposited metal intracellu- larly [10], and as a result the toxicity of metal was reduced so as to provide a condition favorable to the growth and the carbon utilization of the soil microorganisms. The above two reasons might be responsible for the higher Cmic and Cmic/Corg in B soil. There is accumulating evidence that the higher qCO2 is observed in most of metal-contaminated soil, and the qCO2 was about 2.0 times greater in contaminated soil than that in uncon- taminated soil [32,39]. Dai et al. [40] also found that the qCO2 decreased with the alleviation of metal toxicity. The qCO2 in both soils decreased markedly after 12 days of bioremediation (Table 2), which might indicate the Pb toxicity in both soils was alleviated. The qCO2 in A soil was about 1.6 times greater than that in B soil after remediation. Microbial

metabolic efficiency was inhibited in the presence of metals, and more energy was required for microbial synthesis. Microorganisms mainly utilized organic carbon in soil as maintenance energy for growth under metal contamination, so the CO2 release increased and the conversion of organic carbon into biomass decreased. On the contrary, microorganisms could convert substrates into biomass effectively with less or no metal D.-L. Huang etal. / Journal of Hazardous Materials B134 (2006) 268-276 contamination. These findings might explain the change of the qCO2 observed in our study, the similar results were also shown by Dai et al. [40]. The microorganisms (such as fungus and bacteria) with different C/N ratio contributed to the biomass C/N ratio for soil. Generally, there is the dominance of fungi in polluted soils and bacteria in the uncontaminated soils, due to the tolerance of fungi to metal [41]. Because of the difference among the C/N ratios of microorganisms (e.g. fungi, bacteria, etc.) the biomass C/N ratio in soil varied with the unequal growth of different microorganisms. The Cmic/Nmic in the samples displayed a decrease when the Cmic increased (Table 2), which might be because the microorganisms with low C/N ratio increased faster than those with high C/N ratio. Joergensen et al. [42] also confirmed that large microbial biomass C/N ratios were caused by increased fungal to microbial biomass ratios under metal contamination, and they reported that C/N ratio for bacteria was as low as 3.5:1 and fungi, from 10 to 15:1. Although the Cmic/Nmic in B soil decreased more than that in A soil during incubation, our results showed the lower Cmic/Nmic in A soil with higher contamination level and the higher Cmic/Nmic in B soil with lower contamination level on day 60 (Tables 1 and 2), which was opposite to the previous studies and the conclusions we obtained above. The Cmic/Nmic was not suitable to be used as an index for estimating metal toxicity and contamination level in contaminated soil with inoculation of microorganism, which might be because the C mic/Nmic from the inoculated microorganism would affect this estimation. The microbial indices mentioned above just reflect the difference of growth of soil microorganisms and microbial activity under Pb-contamination, but not show the changes in microbialcommunity status. The BIOLOG procedure is a useful method for the characterization of the metabolic capacity of soil micro- bial communities. The differences between the metabolic activities of microbial communities can be assessed by analyzing the utilization of carbon sources in BIOLOG plates [43,44]. The higher K and the lower S were observed in B soil after remediation (Fig. 4), which showed the co-incubation method adopted in our study improved the utilization of carbon sources by soil microflora. It was also observed that there was a significant difference between kinetic parameters K and S fitted to logistic growth curve in A soil and those in B soil, while no significant difference between the AWCD in A soil and that in B soil was found (Fig. 4). The above results indicated the kinetic parameters were more sensitive to the changes of metabolic capacity of microbial community than the AWCD, which was used to evaluate microbial community activity and soil ecological status, in accordance with the previous reports [24].

277

5. Conclusions
In conclusion our results showed that incubating contaminated soil with the inoculated P. chrysosporium, together with the added straws as nutrient, could reduce the active Pb, alleviate the Pb stress, and stabilize the Pb-contaminated soil. In addition, the treatment improved the remediation of the soil in comparison to the controls. All these results might be because the Pb ion was absorbed by the mycelia of P. chrysosporium and chelated by the humus formed in the incubation process. However, further studies are needed to investigate and confirm the immobilization mechanisms of Pb, because the mechanisms were not very certain in this study. The traditional bioremediation technology relying on plants might be difficult to remediate some lean soil contaminated with heavy metal because the soil conditions are unfavorable to the growth of the plants. In addition, when soil microflora had deteriorated, conventional physico-chemical approach could not improve soil microbial activities effectively but only removed metal from contaminated soils. The approach used in our study could not only immobilize Pb ions in soils, but also effectively improve soil microbial activities and the metabolic capacity of microbial community. However, metal ions were just immobilized for complexation and not removed from soil in our study, so the approach needs improvements and deserves further researches.

Acknowledgements
The study was financially supported by the National Natural Science Foundation of China (Nos. 50,179,011 and 70,171,055), Chinese National 863 High-Tech Program (2001AA644020, 2003AA644010 and 2004AA649370), the National Basic Research Program (973 Program) (2005CB724203), Chinese National Natural Foundation for Distinguished Young Scholars Project (No. 50225926), Research Award Program for Outstanding Young Teachers in Higher Education Institutions of Ministry of Education (MOE), P.R.C. 2000, the Doctoral Foundation of MOE of China (No. 20020532017).


相关文章:
更多相关标签: