给排水英文文献翻译
附录一 外文资料翻译 罗马尼亚过去30年污水处理的演变过程： 案例研究勒姆尼库沃尔恰 L.波特尔 摘要 -- 罗马尼亚的污水处理工艺，在过去30年的演变：案例研究勒姆尼库沃尔恰。大多数站在国家层面的水文领域研究都是关于地表水的话题，这些研究分析了水文、水文特征和某些集水区的水的质量。 本文的目的不仅仅介绍城市水文学的相对较新的领域，更准确的说应是城市污水处理工艺。这项研究是基于在2010年上半年从勒姆尼库沃尔恰污水处理厂和一些相关的污水水质指标和治理过程中的意见以及2005年和2011年之间分析得出来的。该研究显示，在罗马尼亚的城市污水处理过程的效率推出的2000年后，新的处理工艺和技术的重要性。 关键词：污水处理厂，主要处理，二级处理，三级处理，效率。 1.介绍 城市污水代表的是一个城市的范围内生活用水，工业用水和雨水，并由运输和疏散出来的城市的污水管网收集的混合物，如果城市污水排放到地表水没有适当的处理，它会成为环境污染的一个重要因素。
城市污水处理的专门设施即污水处理厂，只设在城市和城镇的集中式污水
处理系统。这些污水处理厂的运行原理并没有在过去三十年的变化，目的是为了满足对废水排放到环境中的要求。 一个符合有关地表水和地下水的保护条例的污水处理厂，包括两个不同的技术流，即水流量和污泥流。本研究只考虑了水流，它是一个在过去的三十年已经知道的最显著的进化。对于国家层面来说，到2000年为止，这种以目的来指导的变化可分为两个时期。在过去的十年间，来自罗马尼亚的污水处理过程中经历了与欧洲的立法在水领域遵守的必要性支持的一个重要演变，主要是1991年5月的理事会条例91/271 / EEC关于城市垃圾,水的处理,并在1998年2月由欧盟委员会修改为条例98/15 / EC，在国家立法中政府决议并通过。关于对废水排放到水生环境的条件的规则批准，由政府补充和修订。 所提到的演变决定着在罗马尼亚区域废水显著的改善情况。然而，该国未能达到前两个中期期限的理事会条例271/91/ EEC的实施。 该国的污水管网在2010年12月31日报告管总长是22,196.34公里，连接这个网络之中的人口比例为54.28％，43个超过2000居民的城区的管网连接区域人口超过95％。在同一天，420座污水处理厂投入运作，连接到它们的人口占全国总人口的43.24％，这些污水处理厂中，32个已经连接超过这个居民城区人口的95%。然而，三分之二以上的总产生的废水并没有在被排放到环境中之前作适当的处理。通过比较，从而得知勒姆尼库沃尔恰下水道网络在2010年12月31日的长度是146.3公里以及连接到该网络的人口的百分比为75％。
2.在2000年以前的一段时期内
在2000年之前建成的污水处理厂有通常是由两个处理阶段，即一级和二级
处理工艺流程。一级处理（或机械处理）从水中清除大型的漂浮物，倾析杂质和那些被漂浮或可以处于浮动状态进行运输的物质。在此阶段，应先进行预处理，这个操作是从水中移除某些物质（沙，脂肪），以防破坏二级处理的装置；二级处理（或生物处理）需保证污水溶解度，除去不能在一级处理过程中胶体分散的有机杂质。这些杂质是通过微生物在无害降解产物的培养和在一个新的细胞块进行转换的。通常，微生物的这种培养物分散在处理装置的反应体积，该过程被称为具有活性污泥的生物处理。 这样的技术流从勒姆尼库沃尔恰废水处理厂取代旧的（2009年12月）开始就定义了废水处理。在这段时间内，该污水处理厂在1989年进行了单独的重要调整过程,它的处理能力增加了一倍，达到了1020升/秒。在机械处理的过程中最重要的组成部分（图1）由两个特殊密集屏幕组成用于从水中去除直径大于2厘米的杂质：包括3个沙砾去除通道，一个除脂通道和两个初级沉淀池，能够除去杂质并且每个都能够容纳3000立方米。生物处理过程（图1）的最重要的组件是两个表面 -曝气池，其提供了在工艺流程中微生物使用必要的氧气，每个曝气池都有3000立方米，两个二级沉淀池再除去降解产物和由微生物产生的细胞残留物，每个二沉池的体积达1500立方米。 图1。勒姆尼库沃尔恰在一九八九年 至2009年期间，污水处理厂工艺流程图
3.2000年之后的时期内 一种新的、更高效的污水处理厂的模式，即在水领域符合欧洲法规，BAS在过去十年被罗马尼亚引入。新的污水处理厂是由欧洲资助的，这些项目通常把现有的污水处理厂的升级，而不是建立新的处理厂。该模型保持原有的一级和二级处理，并具有一些改进，主要是辅助特有的早先类型污水站处理的过程，加入三级处理过程，这个处理过程也称为化学或高级处理过程。新过程中从水移除营养物质（化合物氮和磷），并减少微生物的数量。 氮的去除是通过两个独立的过程完成的，即硝化和反硝化。第一个是在需氧过程，第二个要求缺氧条件。它们常常由专门的细菌通过活性污泥法中的水引入。在二次处理用曝气池 - 无论硝化和反硝化可以在表面内完成。磷的去除，通常使用的铁或铝的盐化学沉淀来实现的。在消毒过程中大部分微生物是通过用氯、臭氧或紫外线除去的。 图2。从勒姆尼库沃尔恰污水处理厂开始， 到2010年所述初级处理阶段过程流程图 这样的处理厂规模从2008年至2009年间在勒姆尼库沃尔恰实施的，它开始于2010年1月1日，并从现有设施升级为一个更有效地发挥作用的设施。代表项目“整修勒姆尼库沃尔恰的实施结果"其中就涉及了1082万欧元投资的污水处理厂。这一数额的51.49％，是由欧盟提供的ISPA基金以及另外的48.51％是由欧洲投资银行的贷款。 从勒姆尼库沃尔恰新的污水处理厂1550升/秒的处理能力，其初级处理阶段较旧的污水站（图2）具有重要差异。沙子和油脂的去除由单一结构制成的充气室完成，由于它们的低效率，在初沉池阶段就去除。在下大雨的情况下，这就超过了处理厂的容量水的量，这些水就会被存储在两个8800立方米的总容量池中。如果水量满出后，这些未经任何处理的水就直接排入奥尔特河中。当在污水站中废水的流量低于预定值，在两个容器中累积的水就会被适当的处理。 新的污水处理厂的二级处理阶段(图3)与旧的没有明显区别，唯一区别就是较大的表面 - 曝气池，其每个曝气池容量为11400立方米；并且更重要的是，四种终沉池每个容量都为13000立方米，它又能够同时满足沉淀池的功能，是在旧处理厂的情况下发展的新类型。 目前三级处理阶段是氮和磷的去除（图3）表示。 图3。二级和三级处理工艺流程图 勒姆尼库沃尔恰污水处理厂2010年开始沿用 该除氮的一部分是在富含氧气的上表面 - 曝气池（硝化部分）进行，而另一个是由在其下表面 - 曝气池（反硝化部分）进行；其中，所述溶解的氧被生物体完全使用。除磷是通过化学沉淀进行的，使用氯化铁的盐溶液（FeCl3)一般是用40％氯化铁的浓度，经营污水处理厂的公司希望通过专门的除磷菌提高这个阶段包括水和氯消毒和增加养分去除过程的效率（勒姆尼库沃尔恰增加脱氮除磷工艺的效率，Ediţia revãzutã 03,2009）。
4.处理过程的效率 比较在本文所提出的两个污水处理厂的类型的效率，最好的方式是分析在特定所监视水质指标以及预定时间段的变化，这些都是从污水处理厂里并在化学实验室通过一定的方法所决定的。本研究采用的11项指标的年平均值来自未经处理的水和充分处理的。通过特性的两种类型的水的值之间的差获得每个指标的整体效率。2005年和2009年期间代表勒姆尼库沃尔恰老污水处理厂，而2010年和2011年期间则代表一个新的处理厂。在下面的段落和图像所使用的缩写为：COD - 铬（使用重铬酸钾化学需氧量），COD - 锰（用高锰酸钾化学需氧量）和生化需氧量（生化需氧量）。 新的污水处理厂模式引入技术和设备使得在处理过程显著整体改善，事实可在2010年9项指标值的降低与2009所显示的比较（图4，图5）看出。最重要的差异是化学需氧量的绝对值的发现 - 因为其数值大，通常表征这些指标是铬、锰和COD。在2010年唯一偏高的参数是氯化物和过滤残余物，这一事实表明在这些方向上的新的处理效率已在下降。 两个污水处理厂模式之间的差异更加明显，如果我们分析2005年和2011（图6，图7）之间效率的演变，除氯化物和过滤残余物外，2010年和2011年所有其他指标的效率大幅增加。在一些情况下，这两种类型的处理厂（即间2009和2011）之间的差异是非常高，例如洗涤剂达到70％和铵和硫化物达到约60％，而在其他情况下的值是低得多的（几乎7％pH值），甚至为负（约4％氯化物和过滤残余物）。然而在旧处理厂的情况下，对于大多数指标效率都不超过60％，这种情况在新的处理厂的情况下完全相反，具有高于80％的值，甚至部分指标达到90％。2011年的效率与2010年相比，大多数的指标更高，尽管在第一种情况下，绝对值一般较大，但由于增加了水在处理厂记录的数字。
图4.完全处理水的监测参数（I） 图5.完全处理水的监测参数(II) 图6.处理过程被监控有关参数的总体效率（Ⅰ）图7.处理过程被监控有关参数的总体效率（II） 两个污水处理厂类型的初级处理阶段之间的一个重要区别是有无初级沉
淀池的工艺流程。这一措施可从2005和2009（图8）之间它们的效率转变得到论证。事实表明这些值在所考虑的期间内正连续不断地下降，以至于在2009年降至低于15％。这一下降趋势使可持续旧处理站的整体效率得到增长，尽管二级处理阶段在大多数情况下提高了分析周期值。
图。8污水处理厂的处理阶段的效率 5.结论 罗马尼亚在过去三十年污水处理过程中的演变，实质上也表明地表水和地下水的保护正变得尤为重要。在2000年之后，这种演变的关键时刻通过在水领域中以欧洲法令的国家立法的转换被代表。这一事实决定的许多现有的机械–生物
给水排水专业英语翻译
《给水排水专业英语》译文：（第一课）给水工程我们知道，水的供应对生命的生存至关重要。
人类需要喝水，动物需要喝水，植物也需要喝水。
社会的基本功能需要水：公共卫生设施的冲洗，工业生产过程耗水，电能生产过程的冷却用水。
在这里，我们从两方面讨论水的供给：）1、地下水供给2、地表水供给地下水是通过打井而得到的重要直接供水水源，也是一种重要的间接供水水源，因为地表溪流（或小河）会经常得到地下水的补给。
在靠近地表的通气层中，土壤孔隙内同时包含着空气和水。
这一地层，其厚度在沼泽地可能为零，在山区则可能厚达数百英尺，蕴涵三种类型的水分。
重力水，是在暴雨过后进入较大的土壤孔隙中的水。
毛细水是在毛细作用下进入较小的土壤孔隙中的水，它能够被植物吸收。
吸湿水是在不是最干燥的气候条件下由于分子间引力而被土壤稳定下来的水。
地表通气层的湿气是不能通过凿井方式作为供水水源的。
位于通气层以下的饱和层，土壤孔隙中充满着水，这就是我们通常所说的地下水。
包含大量地下水的地层称为含水层。
通气层和含水层之间的水面称为地下水位或浅层地下水面，地下水静压力与大气压力相等。
含水层可延伸相当深度）, but because the weight of overburden material generally closes pore spaces（但因为地层负荷过重会压缩（封闭、关闭）土壤孔隙，深度超过600m，即2000英寸，就基本找不到地下水了。
能够含水层中自由流出的水量称为单位产水量。
The flow of water out of a soil can be illustrated using Figure 1（土壤中水流如图1所示）. The flow rate must be proportional to the area through which flow occurs times the velocity（流量与流水面积成比例，流经该土壤面积的流量等于面积与速率成的乘积）, orQ=AvWhere（此式中）Q=flow rate , in m3/sec（流量，单位为m3/s）【cubic meter per second】A=area of porous material through which flow occurs, in m2（渗透性土壤的流水断面，单位为m2）v=superficial velocity, in m/sec（表观流速（表面流速），单位为m/s）表观流速当然不是水在土壤中流动的真实速度，因为土壤固体颗粒所占据的体积大大地降低了水流通过的空间。
给排水英文翻译
Removal of inorganic anions from drinking water suppliesby membrane bio/processesSvetlozar Velizarov*,Joa˜o G.Crespo & Maria A.ReisCQFB/REQUIMTE,Department of Chemistry,FCT,Universidade Nova deLisboa,P-2829-516 Caparica,Portugal (*author for correspondence,e-mail: velizarov@dq.fct.unl.pt)Received 29 June 2004; accepted 8 October 2004Key words: Donnan dialysis,drinking water,electrodialysis,inorganic anionic pollutants,integrated processes,membranebioreactors,nanofiltration,ultrafiltration,reverse osmosisAbstractThis paper is designed to provide an overview of the main membrane-assisted processes that can be used for the removal of toxic inorganic anions from drinking water supplies.The emphasis has been placed on integrated process solutions,including the emerging issue of membrane bioreactors.An attempt is made to compare critically recently reported results,reveal the best existing membrane technologies and identify the most promising integrated membrane bio/processes currently being under investigation.Selected examples are discussed in each case with respect to their advantages and limitations compared to conventional methods for removal of anionic pollutants.The use of membranes is particularly attractive for separating ions between two liquid phases (purified and concentrated water streams) because many of the difficulties associated with precipitation,coagulation or adsorption and phase separation can be avoided.Therefore,membrane technologies are already successfully used on large-scale for removal of inorganic anions such as nitrate,fluoride,arsenic species,etc.The concentrated brine discharge and/or treatment,however,can be problematic in many cases.Membrane bioreactors allow for complete depollution but water quality,insufficiently stable process operation,and economical reasons still limit their wider application in drinking water treatment.The development of more efficient membranes,the design of cost-effective operating conditions,especially long-term operations without or with minimal membrane inorganic and/or biological fouling,and reduction of the specific energy consumption requirements are the major challenges.Abbreviations: D – dialysis; DD – Donnan dialysis; DMB – dialysis membrane bioreactor; ED – electrodialysis;IEMB – ion exchange membrane bioreactor; MCL – maximum contaminant level; MF – microfiltration;NF – nanofiltration; RO –reverse osmosis; TOC – total organic carbon; UF – ultrafiltration; USEPA –United States Environmental Protection Agency; WHO – World Health Organization1.IntroductionA number of inorganic anions have been found in potentially harmful concentrations in numerous drinking water sources (DeZuane 1997; Smith et al.2002; Petrovic′ et al.2003).The maximum allowed concentrations of these compounds are generally set by the drinking water quality regulatory standards in the relatively low lg/l)1–mg/l)1 range; therefore,the majority of them can be referred to as micropollutants.The internationally accepted standards and guidelines,regarding the maximum allowed levels of these compounds,are proposed by the World Health Organization (WHO).In addition,the European Union and the US Environmental Protection Agency have issued similar health and environmental standards and a considerable number of regulatory methods have been published worldwide for the analysis of inorganic anions in drinking water (US EPA 1998; Jackson 2001).Table 1 lists the current status for potentially toxic inorganic anions along with some information about the main sources of pollution and the potential healthrisks,associated with their ingestion in drinking water Since there are usually no organoleptic changes in drinking water that can be attributed to the presence of toxic inorganic anions in trace levels,it is rather possible that some of them may remain undetected,thus increasing the possible health risks.An example of such recently found compound is perchlorate,an important ingredient of solid rocket propellants,which may interfere with the ability of the thyroid gland to utilize iodine in hormones production (Richardson 2003; Min etal.2004).Since perchlorate is a serious problem in some US regions,a provisional drinking water goal of 1 lg/l has been suggested (US EPA 2002).A number of inorganic anionic contaminants can be present at the same time in rather different levels (e.g.nitrate and perchlorate),thusleading to the emerging issue of their control and simultaneous removal from drinking water supplies.Finally,water of defined ion composition is required in the manufacturing of a number of foodproducts,pharmaceuticals and in the fresh water fisheries and sea aquariums.Several common treatment technologies are nowadays used for removal of inorganic contaminants from water rge-scale plants usually apply coagulation with aluminium or iron salts followed by filtration but a number of anions (e.g.nitrate) have very little tendency to coordinate with metal ions and low potential for co-precipitation (Duan & Gregory 2003).Smallscale treatment facilities often use ion exchange and/or adsorption due to their ease of handling and compactness;however,regeneration and additional costs,associated with the disposal of the regenerants used,represent serious problems.Moreover,release of undesirable organics (such asstyrene,divinylbenzene,trimethylamine,etc.) from some synthetic resins to the treated water still hinders a larger application of ion exchange for drinking water production (Kapoor & Viraraghavan1997).Membrane separation processes such as reverse osmosis(RO),nanofiltration (NF),ultrafiltration(UF),microfiltration (MF),dialysis (D),Donnan dialysis (DD) and electrodialysis (ED),if properly selected,offer the advantage of producing high quality drinking water.In many cases,one membrane process can be integrated with another to produce water of even higher quality.In these processes,the membrane can be viewed as a barrier between contaminated and purified water streams.The separation of the two streams often allows for operation with no or minimal chemical water pre-treatment,which otherwise can form deleterious by-products (Bergman 1995; Jacangelo et al.1997).However,in physical membrane processes,inorganic anions are not destroyed but normally concentrated and the concentrate disposal can be costly and difficult to permit in many cases; therefore,post-treatment of the concentrate stream or hybrid membrane-assisted technologies capable of converting anionic contaminants to harmless products are highly desirable.Chemical,electrochemical or biochemical treatment processes are able to deal efficiently with anions (Daub et al.1999; Carraro et al.2000;Centi & Perathoner 2003).Among them,biological reduction is especially appropriate since it offers selective removal of the target anion from water due to anoxic bacteria,which under appropriate conditions (pH,oxidation–reduction potential,temperature,etc.) can use anions as electron acceptors and organic (heterotrophic microorganisms) or inorganic (autotrophic microogranisms) compounds as electron donors for their growth.However,the main concern of using bioprocesses is the risk of secondary water pollution by cells,incompletely degraded nutrients and metabolic by-products,which can promote microbial growth in water distribution systems,thus requiring extensive post-treatment in order to produce safe and biologically stable water.These problems can be overcome,or at least reduced,by introducing a membrane unit as a pre- or posttreatment stage in the water production process.When membrane(s) are integrated with an appropriate bioprocess in a single process,the configuration is generally referred to as ‘‘membrane bioreactor’’.In this paper,discussion on useful applications of membrane processes for removal of inorganic anionic contaminants from drinking water,with a special emphasis on the emerging issue of membrane bioreactors,is presented.Since it is rather difficult to select which membrane applications might be referred as specifically anion removal ones (other water constituents are often also removed),the authors’ opinion,in some instances might be subjective.In most cases,only recentexamples are referred to since the articles selected often familiarize the interested reader with the general history of a specific problem.Where it is appropriate,the discussion starts with a given membrane separation used as a single process and then moves on to its integration in membrane bioprocesses.An attempt is made to identify the advantages,limitations and future research needs in each case.Some of the processes discussed at the present time could be considered still far from practical implementation,but nevertheless,they might provide useful information,on which future developments could be based.2.Pressure-driven membrane processes and membrane bioreactorsPressure-driven membrane processes use pressure difference between the water to be treated and a permeate side as the driving force to transport water across the membrane.These membrane processes include RO,NF,UF and MF.The operating trans-membrane pressure ranges vary significantly but are usually: 20–100 bar for RO,5–20 bar for NF,2–5 bar for UF,and 0.1–2 bar for MF.It has to be mentioned that while MF and UF membranes have well-defined porous structure,for RO and NF membranes the term ‘‘pores’’may be better associated with the intramolecular voids within the polymeric matrix. Another possible classification can be based on the molecular mass of the solute to be separated by a given membrane process that is usually of up to 100 Da (RO),100–500 Da (NF) and 500– 10,0000 Da (UF).Obviously,UF and MF are not suitable for the direct removal of inorganic anions from solution; however they can be implemented in hybrid processes,which produce larger aggregates (subsequently filtered) (Han et al.2002; Yoon et al.2003) or in bioprocesses with the purpose to retain microbial cells,using anions as electron acceptors.On the other hand, RO and NF can be used for removal of inorganic anions as single processes.The advantages and limitations of each process will be discussed in the following sections.2.1.Reverse osmosisReverse osmosis is a well-established technology used for many years in water desalination.Reverse osmosis membranes areasymmetric,i.e.,consist of a thin polymer (nowadays,mostly polyamide) layer combined with a porous support to guarantee the membrane mechanical stability.The RO membranes discriminate on the basis of molecular size and due to the dense properties of the separating layer very high (often close to 100%) retention of low-molecular mass compounds and ions (total desalination) can be achieved.Moreover, the process can be easily automated and controlled.The treated water stream,however,may lack the right balance of minerals and has unpleasing taste because of the retention of all ions (Nicoll 2001).Another disadvantage of RO is the high energy consumptionneeded to maintain the required pressure difference.Reverse osmosis membranes are very sensitive to polarization phenomena (ions accumulation) at the membrane surface contacting the concentrate (retentate) side.In addition,the solubility products of some salts in the retentate can be exceeded,forming precipitates(mineral fouling) besides possible biological fouling due to natural organic matter and microogranisms present.The presence of non-toxic components,such as hardness (Ca2+,Mg2+) 2-anions,can interfere with the separation of toxic anionic species and SO4due to problems that these components may cause with water recovery and ionic strength (osmotic pressure) (Ritchie & Bhattacharyya 2002). Therefore,the contaminated water usually requires pre-treatment (other membrane processes like MF and/or UF are nowadays more and more used) before entering the RO modules.Concerning the RO applications in the case of toxic inorganic anions,few studies have been recently performed mostly aiming at the removal of arsenic species and nitrate (Table 2).Brandhuber and Amy (1998) showed that if the main arsenic species are present as As(III) only RO membranes would be effective.In a pilot-scale study,removal efficiencies between 96–99% for As(V) and 46–84% for As (III) have been reported (Ning 2002).Pre-oxidation of As(III) to As(V) could guarantee better removal.It was demonstrated that RO could be effectively applied for removal of nitrate along with water desalination in a rural area (Schoeman &Steyn 2003).The nitrate removal efficiency was close to 98% and,although the total dissolved solids (TDS) in the treated water stream were strongly reduced (from 1292 mg/l in the source water to 24 mg/l in the RO permeate),the authors suggested that the water could be used directly for potable purposes.Preliminary estimates showed that for an approximately 2 m3/h output plant,the capital and operating cost were about USD 29,900 and 0.50/m3,respectively.It can be concluded that RO is a highly efficient process for removal of inorganic anions from drinking water,which guarantees a secure detoxification of the water supply.However,total desalination is undesired due to possible corrosive problems if water hardness is reduced to very low levels.Water with hardness values under 50 mg/l is expectedto be corrosive (lead,copper,iron,zinc,etc.) (DeZuane1997).Therefore,modifications, allowing for selective toxic anions removal along with a sufficient retention of water salinity are required.2.2.NanofiltrationNanofiltration (NF) uses membranes,which can provide selective desalination,and is usually applied to separate multi-valent ions from monovalent ones; however,it is also possible to achieve a certain separation of ions of the same valence by selecting the proper membrane and operating conditions (Lhassani et al.2001).NF membranes are sometimes designated as ‘‘loose’’ RO membranes (Ho & Sirkar 1992),since they provide higher water fluxes at lower trans-membrane pressures.These membranes are usually asymmetric and negatively charged at neutral and alkaline drinking water pH.Therefore,separation of anions is based not only on different rates of their diffusion through the membrane (at low pressure),convection (at high pressure),but also on repulsion (Donnan exclusion) between anions in solution and the surface groups,which is obviously higher for multi-valent anions (Levenstein et al.1996).The advantage of introducing this additional mechanism of ion exclusion (in addition to the size-based exclusion) is that high ion separation degrees (ion rejections) similar to those in RO can be achieved but at higher water fluxes through the membrane.On the other hand,the NF process is much more sensitive than RO to the ionic strength and pH of source water.The membrane surface charge is mainly due to anion adsorption from water rather than to fixed charged groups (as in the case of ion exchange membranes), therefore it depends strongly on bulk anion concentration (Hagmeyer & Gimbel 1998).Furthermore,it changes from negative to zero net charge at the membrane isoelectric point and then to positive at lower pH values (usually <4) due to cation adsorption.This pH dependence can strongly affect the target anion separation. Therefore,the selection of adequate operating conditions is more critical for NF applications. Despite these challenges,a number of studies dealing with the removal of toxic anions from drinking water have been performed (Table 2) and most of them have showed promising results.Experiments with groundwater,to which arsenate As(V) and arseniteAs(III) were added,were performed by Urase et al.(1998),who showed that the As(V) rejection between 90 and 97% from the negatively charged NF membrane used was almost not influenced by water pH; however,the As(III) rejection increased with pH,being 50% at pH 3 and 89% at pH 10.At pH 10,most of arsenite is in a mono-valent anion form,while at low pH the neutral form dominates because the pKa value of arsenite is 9.1.For the same reasons,As(III) was not effectively removed in two other studies (Vrijenhoek & Waypa 2000; Sato et al.2002) while As (V) removal reached 90 and 95%,respectively.It was found that the rejection of As(III) decreased with increasing bulk concentration,an effect that was attributed to its enhanced diffusion and convection through the membrane under such conditions (Vrijenhoek & Waypa 2000).Fluoride removal by NF has also been tested for model water (Lhassani et al.2001) and contaminated groundwater (Cohen & Conrad 1998). In a model study,NF showed potential to selectively separate the following single halide salts: NaF,NaCl,NaI,LiF and LiCl (Diawara et al. 2003).The fluoride selectivity was higher at low fluxes,where the chemical parameters (e.g. hydration energy,partition coefficient) are predominant and larger ions were less retained. Therefore,fluoride,with a hydration energy higher than those of chloride and iodide,was better retained despite the fact that it is the smallest anion.Nitrate removal from drinking water by NF has been the subject of several studies (Table 2). A reasonably high nitrate rejection of 76% was observed by Van der Bruggen et al.(2001),who also performed preliminary cost analysis,showing that for a water treatment capacity of 2000 m3/h the operating cost would be approximately 0.13/m3.However,the hardness rejection was very high,up to 95%,therefore the treated water should be re-mineralised to obtain a hardness of approximately 2 mmol/l.It was suggested that the development of new membranes with the same rejection for neutral molecules but with lower rejection for charged compounds would allow to obtain a permeate with hardness ready for distribution. Nanofiltration is rapidly becoming more and more attractive alternative to the traditional RO water treatment.To a great extent,this is due to the introduction of highly efficient NF membranes and moduleconfigurations,which allow lower investment and operatingcosts.Nowadays,the world’s largest NF plant treating surface water (from the river Oise) is located in France and comprises over 9000 Filmtec NF200 membrane modules able to produce 140,000 m3 of drinking water per day (Nicoll 2001).2.3.Pressure-driven membrane bioreactorsThe concept of integration of biological treatment (aerobic or anoxic) of polluted water with its filtration across a porous membrane,driven by a pressure difference was first applied for the separation of activated sludge by an UF membrane and its recycling to the aeration tank (Smith et al. 1969),and since that time has been exploited extensively in urban and industrial wastewater treatment (Bouillot et al.1990; Brindle & Stephenson 1996; Gander et al.2000; Ben Aim & Semmens 2002; Cicek 2003; Wintgens et al.2003). The polluted water and microbial culture are in the direct contact,while the product water is ‘‘forced’’ through the membrane due to pressure difference (Figure 1).The main advantage is the possibility of achieving high biomass concentrations within the bioreactor; therefore,the plant size can be reduced.As a result of membrane separation,the biomass retention time is independent from the hydraulic retention time,thus allowing slowgrowing microorganisms to be maintained in the bioreactor.This feature is of particular importance for toxic compounds and/or micropollutants, which usually need long periods for their complete biological degradation/transformation to harmless products.Previous articles dealing with pressuredriven membrane bioreactors categorized such systems in terms of the manner in which the membrane module is integrated (e.g.an external membrane module,or membranes directly immersed in the culture medium,a configuration known as an ‘‘submerged membrane bioreactor’’). The latter configuration appears to be less energy consuming and less detrimental for the biomass (Ben Aim & Semmens 2002).Following the successful application of pressure-driven membrane bioreactors in wastewater treatment,the first attempts to extend this approach to the production of drinking water were done in the 1990s (Table 3).Chang et al.(1993) studied denitrification of tap water supplementedwith nitrate salts,which was continuously pumped to an agitated anoxic bioreactor coupled to an external hollow fibre UF membrane (a nominal porosity of 0.01 lm) module,through which the culture medium was recycled.The carbon and phosphorus needed for bacterial growth were provided by ethanol and phosphoric acid,respectively. Efficient denitrification (NO3) and NO2) concentration bellow the respective MCLs) was achieved at a maximum permeate flow rate of more than 100 l/(m2 h) and nitrate removal rate of 11 g/(m2 h).The most serious detected problem was related to the strong decrease of water filtration rate from 120 to 45 l/(m2 h) after 10 days of operation because of membranefouling.Therefore, a backwashing procedure (for 12 seconds every 12 min) had to be implemented,thus increasing the process complexity and energy demands. Furthermore,a secondary pollution of the treated water by 1.5–2.1 mg C/l (as total organic carbon) was found.While this TOC level was much lower than the one in the bioreactor (40–50 mg/l),it was still high enough to promote secondary microbial growth in the distribution systems.Using the same concept but with a plane UF membrane module (cut-off diameter of 200 kDa), Delanghe et al.(1994) also obtained a highly efficient denitrification of tap water,supplemented with nitrate salts,and reached a maximum permeate flow rate of about 21 l/(m2 h) and nitrate removal rate of 3 g/(m2 h).They dealt with the fouling problem by washing weekly the membranes with 3% NaClO solution pumped through the modules.However,the TOC concentrations in the treated water varied between 5–10 mg C/l. Since no ethanol was found,this TOC increase was attributed to low-molecular mass organic compounds, lower than the membrane cut-off diameter. The authors concluded that whatever its origin,this TOC level made the water unsuitable for drinking and needs further treatment.Barreiros et al.(1998),using acetate as the carbon source and electron donor,succeeded to denitrify efficiently naturally contaminated groundwater.They used a hollow fibre polysulfone UF membrane module with a cut-off of 500 kDa and obtained reasonably high water permeate flow rate of about 30 l/(m2 h) at a nitrate removal rate of 4.5 g/(m2 h).TOC values of 1.5–2.0 mg C/l were determined in the treated water,which were similarto those measured by Chang et al.(1993).The dosage of electron donor to the contaminated water has to be carefully controlled in response to the electron acceptor (anion) concentration. On-line monitoring of the target anion concentration in the feed water and adding the carbon source as a function of that concentration by means of an adaptive control system could solve this problem.Such control is important since electron donor limitation would lead to incomplete anion reduction,while overdosing could promote microbial growth in water distribution systems.However,the capability of a pressure-driven membrane bioreactor to provide TOC free water would be still questionable because of possible transport through the membrane of low-molecular organic compounds (other that the carbon source) as noted by Delanghe et al. (1994).Recently,the removal of nitrate from synthetic feed water using a sulfur-based autotrophic denitrification in a pressure-driven membrane bioreactor utilizing a rotating UF membrane disks with a cut-off of 750 kDa was tested (Kimura et al. 2002).While the obtained water permeate flow rate of about 20 l/(m2 h) is within the expected range,this study is interesting due to the use of membranes directly immersed in the bioreactor. Membrane fouling was not severe and cleaning by increasing the disk rotation velocity from 200 to 400 rpm for 15 min every 3 days allowed for a stable process operation for more than 3 months. However,assimilable organic carbon was still detected in the treated water.Fluorescence spectroscopy data showed differences between feed water and filtrate,however,the authors could not identify or characterize the organic matter from the contour plots obtained.Besides their direct application for drinking water production,as discussed above,pressuredriven membrane bioreactors have been occasionally used to treat concentrated water streams, obtained after applying other processes.An advantage of such ‘‘associations’’ is that biological treatment can be applied to solutions whose compositions and temperature can be controlled and is independent of possible changes in the characteristics of water to be treated.Forexample,nitrate-containing brine,obtained after an electrodialysistreatment of groundwater,contaminated by agricultural and farm activities,was studied (Wisniewski et al.2002).A ceramic membrane with an average pore size of 0.05 lm,periodically regenerated by chemical means,was used. An almost total removal of nitrates (99%) was achieved by the mixed culture despite the presence of other ions in relatively high concentrations.The drinking water quality is determined by the process used (electrodialysis in the case studied).Overall,in pressure-driven membrane bioreactors, fouling was found to affect the process performance either due to the deposit of a layer at the membrane surface and/or by partial or complete blockage of the pores.However,fairly efficient solutions (e.g.,periodic membrane backwashing) can be implemented so that relatively high amounts of water may be treated per unit area of membrane.With the reduction of the price of commercial membranes,the process energy consumption is usually the most important economic factor.A general limitation of the pressure-driven membrane bioreactors studied until now is the treated water quality.While contamination of water with microbial cells and biopolymers can be avoided,the retention of ions and low molecular mass compounds (such as some metabolicby-products) by porous membranes is generally insufficient to meet the stringent drinking water criteria; therefore either process modifications or water post-treatment are necessary.3.Dialytic membrane bio/processesDialysis uses a semi-permeable membrane to separatecompounds due to their different rates ofdiffusion in the membrane.Since the dialysis processdriving force is the chemical potential difference(in opposition to pressure difference for RO,NF,UF and MF),one of the main differencesbetween pressure-driven membrane processes anddialyitic processes is that the solvent (water) passesthrough the membrane with more or less selectivesolutes (ions) retention in pressure-drivenprocesses,while the solutes pass through themembrane with more or less selective transfer in dialysis.In practice,dialysis is used for separationof compounds,which differ significantly in size inorder to guarantee a large difference in diffusion rates; the classical example is hemodialysis (artificial kidney) for purifying human blood byremoving small solutes such as urea,chloride,etc.,while retaining large proteins and other componentsin the blood.Therefore,the ability of dialysisto discriminate between different ions in water is limited.Furthermore,the flux of a given compound depends on its concentration gradientthrough the membrane,thus dialysis is characterizedby considerably lower flux rates in comparisonto pressure-driven membrane separations.Therefore,despite its great pharmaceuticalimportance,dialyisis has not been applied as asingle process in drinking water treatment.On theother hand,attempts have been made to separateanaerobic mixed microbial culture from water tobe treated using microporous hydrophobic membranes, based on a dialysis mode of operation,aconfiguration,which might be designated as adialysis membrane bioreactor.Besides the chemical potential (instead of pressure) difference usedas the driving force,another important difference between this type of bioreactor and the pressuredriven membrane bioreactor is that the water to betreated and the microbial culture are separated bythe membrane in different compartments(Figure 1).3.1.Dialysis membrane bioreactorsNitrate removal was studied in a system,in whichmodel water and denitrifying microorganisms were。
给水工程英文文献翻译
附录C：外文文献及其译文外文文献：Removal of Pharmaceuticals during Drinking Water Treatment The elimination of selected pharmaceuticals (bezafibrate, clofibric acid, carbamazepine, diclofenac) during drinking water treatment processes was investigated at lab and pilot scale and in real waterworks. No significant removal of pharmaceuticals was observed in batch experiments with sand under natural aerobic and anoxic conditions, thus indicating low sorption properties and high persistence with nonadapted microorganisms. These results were underscored by the presence of carbamazepine in bankfiltrated water with anaerobic conditions in a waterworks area. Flocculation using iron(III) chloride in lab-scale experiments (Jar test) and investigations in waterworks exhibited no significant elimination of the selected target pharmaceuticals. However, ozonation was in some cases very effective in eliminating these polar compounds. In labscale experiments, 0.5 mg/L ozone was shown to reduce the concentrations of diclofenac and carbamazepine by more than 90%, while bezafibrate was eliminated by 50% with a 1.5 mg/L ozone dose. Clofibric acid was stable even at 3 mg/L ozone. Under waterworks conditions, similar removal efficiencies were observed. In addition to ozonation, filtration with granular activated carbon (GAC) was very effective in removing pharmaceuticals. Except for clofibric acid, GAC in pilot-scale experiments and waterworks provided a major elimination of the pharmaceuticals under investigation.IntroductionIn Germany, some pharmaceuticals are used in quantities of more than 100 t/yr (1). Pharmacokinetic studies exhibit that an appreciable proportion of the administered pharmaceuticals are excreted via feces and urine (2) and thus are present in the domestic wastewater. A further source for the contamination of wastewater is assumed to be the disposal of (expired) medicine via toilets. However, this portion is very difficult to estimate because reliable data are not available. After passing through sewage treatment plants (STPs), pharmaceutical residues enter receiving waters. Point discharges from pharmaceutical manufacturers can also contribute to contamination of rivers and creeks (3). First results concerning environmental occurrence of pharma-ceuticals are reported by Garrison et al. (4) and Hignite and Azarnoff (5), who detected clofibric acid in the lower micrograms per liter range in treated sewage in the United States. Further studies in 1981 in Great Britain revealed that pharmaceuticals are present in rivers up to 1 íg/L (6). On Iona Island (Vancouver, Canada) Rogers et al. (7) identified the two antiphlogistics ibuprofen and naproxen in waste-water. Recent investigations showed the exposure of a wide range of pharmaceuticals from many medicinal classes (e.g,betablockers, sympathomimetics, antiphlogistics, lipid regu-lators, antiepileptics, antibiotics, vasodilators) to rivers and creeks. Reviews from Halling-Sørensen et al. (8), Daughton and Ternes (9), and Jørgensen et al. (10) summarize most of the literature in this new emerging field about the environ-mental relevance of pharmaceuticals.Furthermore, Mohle et al. (11), Alder et al. (12), Ternes et al. (3), and Zuccato et al. (13) have reported the identification of pharmaceuticals in the aquatic environment.Contamination is influenced by the relative portions of raw and treated wastewater (14) such that even small rivers and creeks can be highly contaminated. Groundwater is contaminated with pharmaceuticals primarily by infiltration of surface water containing pharmaceutical residues as well as by leaks in landfill sites and sewer drains. Because of the widespread occurrence of pharmaceuticals in the aquatic environment and sometimes also in the raw water of waterworks, a few cases surfaced where pharmaceuticals were detected in drinking water in the lower nanograms per liter range (15, 16). Although up to now no adverse health effects can be attributed to the consumption of pharmaceuticals at these low concentration levels, based on precautionary principles, drinking water should be free of such anthro-pogenic contaminants.Currently, few papers have been published dealing with the removal of pharmaceuticals in drinking water treatment. Ozonation and especially advanced oxidation processes seem to be very effective in removal of diclofenac, while clofibric acid and ibuprofen were oxidized in lab-scale experiments mainly by ozone/H2O2 as shown by Zwiener and Frimmel (17). Heberer et al. (18) exhibited that reverse osmosis is appropriate to remove a variety of different pharmaceuticals from highly contaminated surface waters.The objective of the work presented here was to study the efficiency of different treatment steps to remove the anti-phlogistic diclofenac, the antiepileptic carbamazepine, and the lipid regulators clofibric acid and bezafibrate during drinking water treatment. Therefore, the primary elimination of the selected pharmaceuticals was investigated under laboratory, pilot, and real waterworks conditions. In addition to processes such as bank filtration and artificial groundwater recharge, widely used techniques for surface water treatment such as activated carbon filtration, ozonation, and floccula-tion were investigated. The monitoring results of two German waterworks are extended by lab- and pilot-scale experiments to obtain more generalized results.Experimental SectionSelected Pharmaceuticals.For all lab- and pilot-scale spiking experiments, four relevant pharmaceuticals (the antiphlo-gistic diclofenac, the antiepileptic carbamazepine, the lipid regulators clofibric acid and bezafibrate) have been selected as target compounds. Their molecular structures are shown in Table 1. These compounds have been chosen because of their predominant occurrence in German feeding waters for waterworks such as rivers, bank filtrates, and ground-water (14, 19). Additionally, the antiepileptic primidone was included in oxidation experiments and a waterworks survey.TABLE1.Selected Target PharmaceuticalsAnalytical Methods.The determination of the pharma-ceuticals was performed using different analytical methods (see Table 2). All methods were based on a solid-phase extraction of the analytes on to RP-C18 or Lichrolute EN material. After solid-phase extraction (SPE) and an elution step with methanol or acetone, the compounds were derivatized using different agents. Either a methylation with diazomethane (20) or a silylation with a mixture of N,O-bis(trimethylsilyl)acetamide (BSA) and 5% trimethylchlo-rosilane (TMCS) (Fa. Fluka, Buchs, Schweiz) were used (60 min at 120 °C) (21). Carbamazepine was determined aftersilylation either by a mixture of MSTFA/TMSI/DTE(N-methyl-N-(trimethylsilyl) trifluoroacetamide/trimethylsilylim-idazol/dithioerytrit; 1000 íL/2 íL/2 íg) (22) or by a mixture of BSA/TMCS. For primidone, an acetylation by acetanhy-dride and ethanolamine was used (22). In all cases, GC-MS was used for the detection of the analytes. Further details of the methods are reported in refs 19-22.All methods enable the precise determination of the target pharmaceuticals in river water and drinking water. An interlaboratory comparison exercise (ICE) between the three participating laboratories at the beginning and the end of the study confirmed the quality of the analytical methods. Groundwater and surface water samples were spiked with the selected pharmaceuticals and analyzed by all three laboratories to confirm the recoveries of the analytes in the respective matrixes. The mean recovery of the spiked concentrations always exceeded 70% through different spiking levels:0.40-0.90 íg/L in surface water and 0.030-0.20 íg/L in drinking water. The relative standard deviations between the three participating laboratories were in general below 25%. Thus, it could be shown that (i) the difference of found concentrations was minor between the threelaboratories and (ii) the spiked concentration could be detected in the groundwater and surface water accuratel.Limits of Quantification (LOQ) and Calibration.The LOQ was calculated according to the German DIN 32645 (23) with a confidence interval of 99% using the standard deviation of a linear regression curve. Calibration ranges from 0.005 to 0.050 íg/L and from 0.05 to 1 íg/L were used with at least seven concentration levels by spiking groundwater. LOQ is another term for limit of determination (LOD) mentioned in DIN 32645. Since the calculated LOQ values were always between the first and the second calibration points, the LOQ used was setas the second lowest calibration point of the linear correlation to ensure a precise quantification. Hence, the LOQ were at least 20 ng/L for diclofenac, carbamazepine, primidone, and clofibric acid and down to 50 ng/L for bezafibrate. However, with a final volume of 100 íL instead of 1 mL, LOQ down to 2 ng/L were achieved for clofibric acid, primidone, diclofenac, and carbamazepine and down to 10 ng/L for bezafibrate. The calibration was performed over the whole procedure after spiking groundwater with the standard mixture of the selected pharmaceuticals. The calculation of the concentrations in native samples was carried out using surrogate standards (see Table 2) and a linear 7-10 point calibration curve.Reference Standards.The reference standards clofibric acid, bezafibrate, carbamazepine, diclofenac,and primidone as well as the surrogate standards meclofenamic acid and 2,3-dichlorophenoxyacetic acid (2,3-D) were purchased from Sigma, Germany; dihydrocarbamazepine was purchased from Alltech, Germany. All standards were dissolved in methanol (1 mg/mL) and diluted with methanol to the final stock solution of 10 íg/mL.Treatment Processes Used in Waterworks.(a) Study of Biodegradation in Batch Experiments with Native Surface Water, Groundwater, and Different Filter Materials. Bio-degradation is one of the crucial factors that determine the elimination of organic compounds during artificial ground-water recharge and bank filtration. To assess the general biodegradability of pharmaceuticals in aquatic environmental matrixes, batch experiments were carried out according to the OECD guidelines for testing chemicals (24). The inoc-culum used consisted of 400 mL of surface water and 400 mL of groundwater mixed with 2 L of MITI basal medium. The MITI basal medium was prepared by mixing 1 L of sterile deionized water with 3 mL of sterilized solutions A-D. Solution A was a solution of 21.75 g of K2HPO4, 8.5 g of KH2PO4, 44.6 g of Na2HPO4â12H2O, and 1.7 g of NH4Cl in 1000 mL of deionized water at pH 7.2. Solutions B-D were solutions of 22.5 g of MgSO4â7H2O, 27.5 g of CaCl2, and 0.25 g of FeCl3, respectively, in 1000 mL of deionized water. The groundwater was taken from a German water catchment area with artificial groundwater recharge using slow sand filtration and bank filtration. The individual concentrations of bezafibrate, carbamazepine, clofibric acid, diclofenac, and ibuprofen were in the batch experiments adjusted to 0.1 and 100 íg/L. The batch experiments were exposed to either individual or a mixture of the selected pharmaceuticals. In stock solutions with ethanol, the concentrations of the tested pharmaceuticals were 0.5 mg/mL or 0.5 íg/mL, respectively. After being diluted (480 íL of stock solution in 2.4 L of culture solution), the concentration of ethanol in batch cultures was 0.02% (v:v). The cultures were always incubated in the dark for 28 d at 14 °C (in situ temperature). For anoxic conditions, 25 mg/L nitrate was added as an alternative electron acceptor. The bottles used were gastight. For aerobic sorption experi-ments, 400 g of sand or 400 g of gravel taken from the underground of a groundwater catchment area was used as inocculum and mixed with 2 L of MITI basal medium (solid phase/liquid phase ) 1:5). Sand that is also used for the slow sand filters of a waterworks consists of a mean grain size range of 0.2-0.6 mm. This filter material showed a moderate permeability with a K f coefficient of 4.3 10-4 m/s. The gravel (natural aquifer sediment) was very heterogeneous with a predominant fraction of 2-10 mm grain size and a K f coefficient of 2.9 10-3m/s. Sterile controls (sterilization for 1 h) were prepared to differentiate between sorption and microbial degradation. The sand contains 3.2 mg/g iron and 0.056 mg/g manganese. Coatings with iron and manganese hydroxides were detected in the gravel but were not quanti-fied.Esterase activities were measured to control the physi-ological status of microbial communities during the incuba-tion of batch cultures. The hydrolysis of fluorescein diacetate (FDA) by esterase enzymes was determined according to the procedure of Schnu¨rer andRosswall (25). A 20-íL volume of FDA solution (20 mg/10 mL acetone, stored at -18 °C) was mixed with 3 mL of sample and 0.5 mL of HEPES buffer (0.1 M N-2-hydroxyethylpiperazine -N¢-2-ethansulfonic acid so-dium salt in deionized water, adjusted to pH 7.5; Merck). After being incubated (sterile conditions, 90 min at 20 °C, darkness), the fluorescein formation was immediately mea-sured with a Perkin-Elmer fluorescence spectrometer LC (excitation at 480 nm, emission at 505 nm).(b) Flocculation.For flocculation experiments in lab-scale experiments, a noncontinual procedure, the so-called “Jar test”, was performed. Spiking concentrations, stirring velocity, and reaction times were selected according to parameters of the two waterworks monitored in parallel. The lab device used consists of glass beakers (v) 2 L) with stator, a stirrer with standardized stirrer geometry, and defined submerged stirring depths. The stirring velocity was adjusted according to the mean velocity gradient (G value), which is proportional to the introduced energy and thus to the aggregation of colloids (26). Under stirring (rpm: 400 min-1), 0.1 mL of iron(III) chloride solution (40%) was added to 1.8 L of raw water (spiked with 1 ig/L pharmaceuticals). After a stirring time of 1 min, pH 7.5 was attained by adding Ca(OH)2 (1 mol/L). Then, the aggregation to microflocs was achieved by stirring slowly for 20 min under 30 min-1. After sedimentation for 20 min, a sample was taken from under the water surface,and the turbidity was measured. These measurements showed that the turbidity was always below 1.5 turbidity units of formazine (TU/F).(c) Activated Carbon Adsorption.Adsorption Isotherms.For the determination of the adsorption isotherms, the following parameters have been used: (i) 200 mL of deionized water or groundwater spiked with initial concentrations of 100 íg/L of the pharmaceuticals under investigation, (ii) pulverized granular activated carbon based on coal, (iii) quantities of activated carbon varied to achieve a final concentration of the pharmaceuticals in the solution that is at least 2 orders of magnitudes smaller than the initial one, (iv) small portions of activated carbon (<0.2 g/L) added as suspension, (v) batches with activated carbon tumbled in250-mL flasks for 24 h, (vi) finally all samples were filtered with 0.45-ím polycarbonate filter and analyzed according to the analytical method described before. Evaluation of the isotherms was performed in double logarithmic scale ac-cording to Freundlich (27, 28). For a single compound, the Freundlich equation q ) Kc n describes the relation between the loading q of the activated carbon and the equilibrium concentration c in the solution. K and n denote the Freundlich parameters.Operation of a Granulated Activated Carbon(GAC) Ad-sorber in Pilot Scale.A pilot plexiglass filter was operated in down flow mode to investigate the removal of the selected pharmaceuticals by GAC filtration. The empty bed contact time was about 10 min with a flow velocity of 10 m/h. The filter was filled with fresh granular carbon based on coal, which is often used in drinking water facilities. The filter was operated with groundwater from a waterworks, which was before aerated and filtered to remove iron precipitations. The influent was spiked with bezafibrate, carbamazepine, diclofenac, and clofibric acid. The pilot filter was operated for nearly 9 months. In intervals of 14 d, the concentrations of the pharmaceuticals were analyzed in the filter influent, at five different heights and in the final filter effluent at a bed depth of about 160 cm. The mean influent concentrations of the pharmaceuticals were 1.8 íg/L for clofibric acid, 1.0 íg/L for carbamazepine, 0.26íg/L for bezafibrate and 0.04íg/L for diclofenac. The different spiked concentrations were due to the limited solubility of the target compounds in the feeding water.(d) Ozonation.In a lab-scale device, water was ozonated in 2-L glass bottles by bubbling ozone through the samples in order to simulate real waterworks conditions. By varying the bubbling time, definite ozone doses in the range of 0.5-3.0 mg/L were introduced into the water. The water was continuously stirred at 900 rpm min-1. After a reaction time of 20 min, the remaining ozone was quenched by adding sufficient sodium thiosulfate solution (c ) 2.2 g/L) to the sample. To determine the transferred ozone doses as a function of the bubbling time, Milli-Q water was ozonated, and the dissolved ozone was measured (external calibration of the ozone doses) according to DIN 38408 using N,N-diethyl-p-phenylendiamine (DPD) purchased from Sigma, Germany (29). The transferred ozone doses through the system into Milli-Q water was further confirmed by the indigo method (30). Flocculated water of a waterworks was spiked with the selected pharmaceuticals (dissolved in 50 íL of methanol) prior to ozonation. Afterwards the ozone was bubbled through the spiked water sample for specific times corresponding to desired ozone doses. The half-life of ozone in the post-flocculated water was approximately 12 min.Sampling Procedure.Water samples were collected in brown glass bottles that had been prewashed with successive rinses of Milli-Q water and acetone and were dried for 8 h at 250 °C. Samples were either extracted immediately or stored at 4 °C for a maximum period of 3 d.Grab samples of the waterworks were taken before and after crucial treatment processes of two German waterworks with different treatment trains. All cooled water samples (4 °C) were analyzed as soon as possible (latest after 3 d).(e) Treatment Trains of the Selected Waterworks.The following treatment processes were applied in the two waterworks selected in the current study.Waterworks I (WW-I).Pre-ozonation (ozone dose: 0.7-1.0 mg/L; contact time: ca. 3 min), flocculation with iron(III) chloride, main ozonation (ozone dose: 1.0-1.5 mg/L; contact time: ca. 10 min), multiple layer filter, and a final GAC filtration.Waterworks II (WW-II).Sedimentation, flocculation with FeCl3/CaOH2, GAC filtration, underground passage, bank filtration, and slow sand filtration.Results and DiscussionStudy of Biodegradation in Batch Experiments with Native Surface Water, Groundwater, and Filter Materials.Experi-ments with batch cultures could provide the first clues on the general potential for biodegradation of pharmaceuticals under different environmental conditions. The relative concentrations (C/C0) of the spiked pharmaceuticals in the batch experiments with surface water and groundwater were nearly constant during the whole exposure time of 28 d (Table 3). All variations of elimination rates were within the relative standard deviation (RSD), which was between 6 and 39%.Thus, it can be ruled out that significant sorption effects and biodegradation occurred in the waters and materials used under anoxic and aerobic conditions. These results suggest that the sorption properties of the selected phar-maceuticals can be expected to be low and that their persistence should be relatively high under real conditions such as slow sand filtration or subsoil passage. However, in complex habitats, the bioavailability and the sorption behavior are determined by various biotic and abiotic parameters that were not simulated in the described batch cultures. Parameters such as the species and physiological status of occurringmicroorganisms, the percentage of humic substances, percentage of iron and manganese hydroxides, pH, etc. can differ significantly according to the actual field conditions. The standardized test used according to the OECD guidelines (24), delivers comparable results for the biode-gradability of substances but cannot be transferred to all natural conditions and account for the various parameters. Therefore, on the basis of the described results, (bio)-degradation or sorption of the selected pharmaceuticals under field conditions cannot be ruled out in general, but they should be relatively low. Sorption of the selected pharmaceuticals on iron hydroxides seems to be insignificant since in the flocculation experiments with precipitated iron hydroxides no reduction of the spiked concentrations was found (see flocculation section below). Furthermore, it was observed that the established microbial activity in the test system was high enough for degradation of dissolved organic matter (DOC) and could not be inhibited by the spiked pharmaceuticals as it can be seen by the esterase activity (Figure 1).Removal after Flocculation with Iron(III) Chloride.Floc-culation in lab-scale (Jar test) with iron(III) chloride exhibited no significant elimination of the pharmaceuticals from raw water. The relative concentration levels (C/C0) after floc-culation were 96 ( 11% for diclofenac, 87 ( 10% for clofibric acid, 111 ( 15% for bezafibrate, 87 ( 12% for carbamazepine, and 110 ( 14% for primidone. Thus, c/c0 of the spiked compounds varied without exception within the RSD. The transference of these results from lab-scale to waterworks conditions was shown by a monitoring of up-scaled floc-culation processes in two waterworks (WW-I, WW-II; see section below: behavior in waterworks) yielding similar results.Activated Carbon Adsorption.Adsorption Isotherms.The assessment of the adsorption properties of single compounds onto activated carbon is often performed by recording adsorption isotherms. Freundlich adsorption isotherms with fresh activated carbon were performed for each of the four selected pharmaceuticals. The isotherms are given in Figure 2. Bezafibrate, carbamazepine, and diclofenac exhibited over the whole concentration range (0.1-100 íg/L) a higher activated carbon loading q than did clofibric acid. Hence, clofibric acid has the lowest sorption affinity on activated carbon. In addition to the selected pharmaceuticals, the isotherm of tetrachloroethene is shown in Figure 2. Tetra-chloroethene was used because its removal by adsorption onto activated carbon in full-scale treatment plants is known to be efficient (31). In a concentration range below 10 íg/L, the isotherms of the pharmaceuticals selected exhibited higher loads on carbon as compared to tetrachloroethene. Thus, it can be concluded that the four selected pharma-ceuticals can be removed efficiently under real conditions by activated carbon filtration in waterworks.Nevertheless, sorption efficiencies are always relying on the competition with other occurring organic compounds. As expected, the adsorption capacity for the pharmaceuticals is lower on activated carbon if other compounds such as natural organic substances compete for the adsorption sites. That can be underscored by a comparison of the Freundlich parameters for the adsorption with deionized water and with natural groundwater (DOC ) 2.0 mg/L; SAC at 254 nm ) 5.8 m-1) given in Table 4. The shift toward lower K values is equivalent to a lower sorption capacity. Especially for clofibric acid the slope of the isotherm (n value) is relatively high in groundwater, which can be interpreted as a low adsorption capacity in the low concentration range. On the basis of the isotherms with natural groundwater, it can be expected that the capacity reduction of activated carbon might be signifi-cant due to competitive adsorption of natural groundwater constituents. Hence, the adsorption capacity of the activated carbon in a fixed bed adsorber in waterworks is expected to be lower for pharmaceuticals than in the isotherm experi-ments performed with deionized water.GAC Filtration in Pilot Scale.In pilot-scale experiments,an activated carbon adsorber filled with activated carbon was operated according to the previous description. The breakthrough curves in different filter bed depths of about 80 cm and 160 cm (end of filter) are shown in Figures 3 and 4. These results coincide very well with the data of the isotherm tests listed in Table 4. Carbamazepine showed the highest adsorption capacity of the selected pharma-ceuticals and can be removed at a specific throughput of about 50 m3/kg in a carbon layer of 80 cm and more than 70 m3/kg in a layer of 160 cm even at a relatively high initial concentration of about 1 íg/L. Clofibric acid, with an initial concentration of about 1,8 íg/L, showed a significantly lower adsorption capacity in the isotherm test and in the pilot-scale experiment. An initial breakthrough of clofibric acid could be observed at a height of 80 and 160 cm at a specific throughput of 10 and 17 m3/kg, respectively. Although lower adsorption capacities in the isotherm test are observed for bezafibrate and diclofenac as compared to carbamazepine, both compounds were removed in a bed depth of 160 cm to a specific throughput of at least 70 m3/kg. The differences between the results obtained in isotherm and the pilot plant experiments might be influenced by the lower initial con-centrations applied in the pilot plant experiments Ozonation.For lab-scale ozonation experiments, floc-culated WW-II water was used. The DOC of the flocculated water was 1.3 mg/L, the pH was 7.8, alkalinity was 2 mmol/L, and temperature was 23°C. The initial concentration of the pharmaceuticals under investigation was 1 íg/L. The ef-ficiency of the ozonationprocess for the removal of the pharmaceuticals turned out to be very product specific. At a small ozone dose of 0.5 mg/L, the concentrations of diclofenac and carbamazepine were reduced by more than 97% while clofibric acid decreased by only 10-15% for the same ozone dose (Figure 5). Even extremely high ozone doses up to 2.5-3.0 mg/L led to a reduction of e40% for clofibric acid. Primidone and bezafibrate were reduced by 50% at ozone concentrations of about 1.0 and 1.5 mg/L, respectively. While applying 3.0 mg/L ozone, still 10% of primidone and 20% of bezafibrate remained. Because of the presence of methanol (used for dissolving the spiked pharmaceuticals), ozone was partly transformed into OH radicals. Thus, the direct ozone reaction was probably underestimated, and the oxidation efficiency under waterworks conditions should be even slightly higher than found in lab scale. Although we did no additional work to elucidate the reactivity of the selected pharmaceuticals with ozone or OH radicals, we can rational-ize these observations based on the chemical structures (Table 1). The reactivity of diclofenac and carbamazepine with ozone is expected to be very high. Rate constants k O3 > 105 M-1 s-1can be expected for deprotonated secondary aromatic amines (diclofenac) and molecules containing nonaromatic double bonds (carbamazepine) (32, 33). For diclofenac, a main oxidation product was detected with a mass spectrum showing an increase of the molecular weight of 16 amu, which is an evidence for substitution of a hydrogen by a hydroxy moiety. A hydroxylation of the secondary amino group is likely but has to be confirmed (e.g., by NMR). Because of missing active sites susceptible to ozone attack (34), reactions of ozone with clofibric acid are expected to be very slow. Thus, OH radical reactions should be predominant with k OH 5 109 M-1 s-1(35). Considering the OH radical activity taken from the prediction for clofibric acid, ozone rate constants for bezafibrate and primidone should result in the middle range (k O3 102-103 M-1 s-1). The reactivity of these pharmaceuticals with ozone can be based on their reactive mono- and disubstituted benzene rings (32). It has to be noted that in the current study only the primary target degradation was investigated, thus further research is es-sential to identify and confirm the structures of metabolites formed by ozonation and to clarify the kinetic behavior.。
建筑给排水英文文章
建筑给排水英文文章Building Plumbing and DrainageBuilding plumbing and drainage systems are an essential part of any construction project. These systems provide a safe and sanitary means of disposing of waste and sewage, as well as supplying clean water for drinking and other purposes. Proper design, installation, and maintenance of plumbing and drainage systems are crucial to ensuring the health and safety of building occupants.The plumbing system in a building involves the distribution of water throughout the structure, as well as the collection and disposal of waste and sewage. Water supply pipes bring clean water into the building, while drain pipes carry waste and sewage away. Plumbing fixtures such as toilets, sinks, and showers are connected to these pipes to provide water for various purposes.The drainage system in a building is responsible for safely removing waste and sewage from the building and transporting it to a treatment facility. This system includes the main sewer line, which connects the building to the public sewer system, as well as various pipes and fittings that carry waste from individual plumbing fixtures to the main sewer line.When designing building plumbing and drainage systems, it is important to consider factors such as water pressure, flow rate, and temperature, as well as the type of building and its intended use. The materials used in these systems should be durable, corrosion-resistant, and able to withstand the demands of daily use.Proper maintenance of building plumbing and drainage systems is essential to ensure their longevity and prevent costly repairs. Regular inspection and cleaning of pipes and fixtures can help prevent clogs and blockages, while timely repairs can address any leaks or other issues that may arise. In summary, building plumbing and drainage systems are critical components of any construction project. Proper design, installation, and maintenance of these systems are crucial to ensuring the health and safety of building occupants, as well as the longevity and efficiency of the building itself.。
给水排水专业英语翻译(全)
《给水排水专业英语》译文：（第一课）给水工程我们知道，水的供应对生命的生存至关重要。
人类需要喝水，动物需要喝水，植物也需要喝水。
社会的基本功能需要水：公共卫生设施的冲洗，工业生产过程耗水，电能生产过程的冷却用水。
在这里，我们从两方面讨论水的供给：）1、地下水供给2、地表水供给地下水是通过打井而得到的重要直接供水水源，也是一种重要的间接供水水源，因为地表溪流（或小河）会经常得到地下水的补给。
在靠近地表的通气层中，土壤孔隙内同时包含着空气和水。
这一地层，其厚度在沼泽地可能为零，在山区则可能厚达数百英尺，蕴涵三种类型的水分。
重力水，是在暴雨过后进入较大的土壤孔隙中的水。
毛细水是在毛细作用下进入较小的土壤孔隙中的水，它能够被植物吸收。
吸湿水是在不是最干燥的气候条件下由于分子间引力而被土壤稳定下来的水。
地表通气层的湿气是不能通过凿井方式作为供水水源的。
位于通气层以下的饱和层，土壤孔隙中充满着水，这就是我们通常所说的地下水。
包含大量地下水的地层称为含水层。
通气层和含水层之间的水面称为地下水位或浅层地下水面，地下水静压力与大气压力相等。
含水层可延伸相当深度）, but because the weight of overburden material generally closes pore spaces（但因为地层负荷过重会压缩（封闭、关闭）土壤孔隙，深度超过600m，即2000英寸，就基本找不到地下水了。
能够含水层中自由流出的水量称为单位产水量。
The flow of water out of a soil can be illustrated using Figure 1（土壤中水流如图1所示）. The flow rate must be proportional to the area through which flow occurs times the velocity（流量与流水面积成比例，流经该土壤面积的流量等于面积与速率成的乘积）, orQ=AvWhere（此式中）Q=flow rate , in m3/sec（流量，单位为m3/s）【cubic meter per second】A=area of porous material through which flow occurs, in m2（渗透性土壤的流水断面，单位为m2）v=superficial velocity, in m/sec（表观流速（表面流速），单位为m/s）表观流速当然不是水在土壤中流动的真实速度，因为土壤固体颗粒所占据的体积大大地降低了水流通过的空间。
英文文献 给排水专业
Research on Facilitation of Biodegradation for Azo Dye Wastewater by Bioelectrochemical TechnologyZHAO Yuhua，CANG Xiaoyi, JIN Decai, DONG RuijiaoSchool of Municipal and Environment Engineering Shenyang Jianzhu University Shenyang，China 110168 zyh088@, 666xiaoyi@,kingdecai123@, dongruijiao@Abstract—Active brilliant red X-3B is a kind of azo dye which is difficult to biodegrade. The wastewater with azo dye is of chromaticity depth, high content of organic compounds, water quality changing great, and seriously impacts environment. Bioelectrochemical hydrolysis coupled with biological contact oxidation (BEH-BCO) was used to treat azo dye active brilliant red X-3B simulation wastewater. In this experiment, synergy of micro-electrolysis and biological hydrolysis was used to improve the efficiency of hydrolysis reactor, and then improve the biodegradation of azo dye wastewater. In the experiment, the HRT of Hydrolysis reactor kept running 12h, and the HRT of biological contact oxidation reactor 7.95h. The electric current densities used in the experiment were 0.024, 0.048, 0.071mA/cm2. This experiment was compared with the biological hydrolysis and biological contact oxidation (BH-BCO), the single biological treatment. Experiment results showed that, the removal effect of active brilliant red X-3B by bioelectrochemical technology was very good in heavy dye mass concentration in raw water (concentration was 50mg/L); and in the range of current density used in test, the treatment effect was increased with the increase of electric current density; when electric current density was 0.071mA/cm2, the average removal ratio of dye mass concentration, colority, CODCr, and NH3-N reached 98.77%, 91.39%, 69.98%, and 90.41% by bioelectrochemical technology respectively, and 9.65%, 18.21%, 31.32%, and 85.69% respectively by the single biological method. There are some reasons for the results. The first reason is that the dye mass concentration of wastewater in raw water was too high, single biological hydrolysis was difficult [1]. The second reason is that, from the measure results of oxidation reduction potential, the hydrolysis reactor was in anaerobic condition in the experimental process, which produced inhibitory environment to bacteria. The analysis of UV-visible absorption spectrum of the influent and effluent from each reactor indicated that the molecular structure of active brilliant red X-3B was destroyed by bioelectrochemical technology and turned into readily biodegradable small molecular organic compound, but it changed little by biological treatment. The measure results of redox potential showed that, the redox potential of mixed liquor in BEH was about -200mV, which is within the range of azo compound redox required standards (-180mV ～ -430mV); the redox potential of mixed liquor in BH was about -152mV, which is not in the required range, and is also not in the range of hydrolysis reactor working normally (about 0mV). The results of measured azoreductase activity show that the azoreductase activity of the single biological hydrolysis was 1.68 mg/L·h, the azoreductase activity of bioelectrochemical hydrolysis was 55.33 mg/L·h. In a word, bioelectrochemical technology could promote the activity of hydrolysis microbe greatly, and obviously improve the decolorization effect of active brilliant red X-3B wastwater. It plays a great role in promoting biodegradation of azo dye wastewater. Keywords-azo dye wastewater; bioelectrochemical; hydrolysis; contact oxidation; active brilliant red X-3BI.INTRODUCTIONAzo dye is widely used in the trades of printing, food, cosmetics, and so on [2]. Azo dye wastewater with huge volume and extensive distribution, changing water quality greatly, high concentration of toxic organic compounds, heavy colority, and complicated biodegradation, is one of the intractable industrial wastewaters. The routine methods to treat dye wastewater include physical method, chemical method, biological method, electrochemical method, and etc. These methods exist obviously disadvantages when they are used [3]. Bioelectrochemical technology is a method that electrochemical reaction and biochemical reaction are set off in a same reactor. It could make electrochemical reaction and microbial reaction complementary and enhance each other, and improve the efficiency of treating wastewater, and reduce the equipment initial investment [4]. In this test, bioelectrochemical technology was applied to treat active brilliant red X-3B that is in a simulated dye wastewater, researched the strengthening effect of bioelectrochemical technology by the contrast experiments between BEH-BCO process and BH-BCO process, and explored the biodegradation mechanism of active brilliant red X-3B. II. TESTER, MATERIALS AND METHODS A. Tester and methods The schematic of the experiment system was showed in Fig.1. This experiment system was divided into two parts, A and B. Part A was BEH-BCO process. In this process, iron sheet (100cm×50cm) in hydrolysis reactor is as the anode, and graphite column (diameter 4cm, high 100cm) as the cathode. Part B was BH-BCO process.This project is supported by Municipal & Environmental Engineering Key Laboratory Open Foundation of Colleges and Universities in Liaoning Province (No.SZ-200901) and Science and Technology Foundation of Ministry of Housing and Urban-Rural Development (No.2010-K7-14).At the bottom of hydrolysis reactor, pulse current which is provided by air compressor play a part in stir mainly, and shall not exceed the limit of dissolved oxygen (DO) in hydrolysis reactor. Continuous aeration was offered to the contact oxidation reactor. Insides of hydrolysis reactor and contact oxidation reactor placed the combination filling (specific surface area 1400-2500 m2/m3). Both BEH-BCO process and BH-BCO process had the same design parameters. Effective volume of BEH reactor was 24L. The hydraulic residence time of BEH was 12h. Useful volume of BCO reactor was 15.9L. The hydraulic residence time was 7.95h. Flow rate was 2L/h. Because the hydrolysis reactor was cylindrical, electric current density distributed uneven. This paper used average electric current density (center plane interfaced between two electrodes with 50% useful volume each).+solution. A standard curve covered colority from 10° to 100° was drawn. The absorbency of the sample which had been centrifugated was measured at wavelength of 350nm. Colority was obtained by calculation [5]. Azoreductase activity was measured by TTC deoxidation method [6]. III. RESULTS AND DISCUSSION The test device adopted domestic sewage inoculation sludge and hanged membrane for 41 days. After the start-up of the device, the results from the contrastively experiments between BEH-BCO process and BH-BCO process to treat active brilliant red X-3B wastewater in electric current density 0.071, 0.048, 0.020 mA/cm2 were showed in figure 2- figure 5. A. Azo dye concentration variety Figure 2 shows the variety of azo dye mass concentration of inflow and effluent to and from every reactor in different electric current density. It depicts that, comparing to the single biological method, treating active brilliant red X-3B by BEH-BCO process has better efficiency. The azo dye concentration of effluent from BEH-BCO process increased with the electric current density declined. When the electric current density was 0.071mA/cm2, the average removal ratio of dye concentration by BEH-BCO process was 98.77%. In BH-BCO process, because of the dye concentration of inflow was too high and the DO in hydrolysis reactor was lower, the bacteria in hydrolysis reactor were restrained. As a result, the average removal ratio of dye concentration was 9.65%.A.BEH-BCO B.BH-BCO 1.Wastewater tank 2..Dosing pump 3.Power 4. BEH reactor 5. Graphite electrode 6 .Packing 7. Iron electrode 8. BCO reactor 9. Aerated conduit 10. Effluent 11. Air compressor 12. Air distributor 13. Time controller 14. Disposed sludge 15. BH reactorFigure 1. Schematic of the experiment apparatusB. Azo dye wastewater Component of simulated azo dye wastewater for experiment is showed in Table I.TABLE ICODCr (mg/L)EXPERIMENTAL WASTEWATER COMPONENTNH3-N/ (mg/L) active brilliant red X-3B/(mg/L) colority pHFigure 2.Variety of Azo dye concentration 122.84-206.41 6.3-7.145.5-128.62.06-6.6240.88-66.26B. Colority varietyC. Analysis method CODCr was measured by fast digestion spectrophotometric method. NH3-N was measured by Nessler's reagent spectrophotometric method. The azo dye concentration was measured at wavelength of 540nm by spectrophotometric method. The sample of the wastewater was centrifugated by a centrifuge at a speed of 4000 revolutions per minute about 10 min. Colority was measured by spectrophotometric method. The dilute H2SO4 solution (ca. 0.02 mol/L) made up of K2Cr2O7 and CoSO4 was used as standard solution for colority measurement. The wavelength of measurement was defined at 350 nm, which was the maximum absorbency of the standardFigure 3. Variety of colorityFigure 3 was the variety of colority of inflow and effluent to and from every reactor in different electric current density. It showed that when electric current density was 0.048 and 0.071mA/cm2, bioelectrochemical technology had a good effect on decolorization to active brilliant red X-3B wastewater. When the electric current density was 0.071mA/cm2, the average decolorization rate reached 91.39%. Biological method had a low decolorization effect to active brilliant red X-3B wastewater. The average decolorization rate was only 18.21%. When the electric current density was 0.020mA/cm2, the colority of the effluent from BCH reactor was higher than the inflow to BCH reactor. The reasons were as follows: on the one hand, when the electric current density was lower, the loose biofilm and the free bacterium in reactor made colority increase; on the other hand, because of the lower electric current density, the OH—produced in cathode was too less, but the Fe3+ produced in anode was excessive, the color of Fe3+ made colority increase. C. Organics varietywhich indicated that hydrolysis process was more efficiency. When electric current density was 0.071mA/cm2, the average removal ratio of NH3-N was 90.41% by bioelectrochemical technology, the average removal ratio of NH3-N was 85.69% by biological method.Figure 5. Variety of ammonia nitrogenousIV.RESEARCH OF PRINCIPLEA. UV-visible absorption spectrum Analysis Azo dye’s active brilliant red X-3B has absorption peaks in the wavelength of 280 nm, 320 nm and 540 nm respectively. The absorption peaks 540 nm in visible light was caused by the n→π system which connected benzene ring to naphthalene ring by azo double-bond. In ultraviolet, the absorption peaks in 280nm, 320nm wavelengths were caused by benzene ring, naphthalene ring, dichloro-methoxy-triazine and so on [7]. When electric current density was 0.071mA/cm2， The analysis of UV-visible absorption spectrum for the influent and the effluent is shown in figure 7.4.0Figure 4. Variety of organics3.5Absorbance(AU)Raw water By BEH By BCO(BEH-BCO) By BH By BCOFigure 4 shows the variety of organics of inflow and effluent to and from every reactor in different electric current density. It depicts that along with the electric current density decreased, the CODCr of effluent from BEH increased. When electric current densities were 0.071, 0.048, 0.020mA/cm2, the average removal ratio of CODCr was 50.08%, 30.11%, 24.71%, respectively. When electric current density was 0.071mA/cm2, the average removal ratio of CODCr was 69.98% by bioelectrochemical technology, and the average total removal ratio of CODCr was 31.32% by biological method. D. Ammonia nitrogenous variety Figure 5 shows the variety of ammonia nitrogenous of inflow and effluent to and from every reactor in different electric current density. As shown in Figure 5, the NH3-N of the effluents from BEH reactor and BH reactor are all higher than raw water. It was because azo double bonds of dye molecule was broken in hydrolysis process, became small molecule organics. Benzenering, naphthalene nucleus or dichlotriazine active group of the dye molecule was broken and ammonia nitrogenous was liberated. So ammonia nitrogenous concentration in the effluent was more than that in the influent,3.0 2.5 2.0 1.5 1.0 0.5 0.0 200 250 300 350 400 450 500Wavelength(nm)550 600650 700Figure 7. UV-visible absorption spectrum of azo dye wastewaterAs shown in figure 7. After the treatment of biological method, the absorption peaks were not changed obviously. It showed that the structure of active brilliant red X-3B was not changed. But, after the treatment by BEH-BCO process, the absorption peaks decreased obviously, and the absorption peaks at 540nm was almost zero. Consider the molecular structure of active brilliant red X-3B, the absorption peaks nearby at 280nm and 320nm are decreased, which explained that benzene ring, naphthalene ring, dichloro methoxy triazine were degraded, unsaturated ring was opened. The absorptionpeaks at the wavelength of 540nm declined obviously, it explained that azo double-bond was opened. But at 215nm, the absorption peaks rose, It explained that there were some aromatic ring compounds in the wastewater [8]. B. Redox potential analysis In order to research the redox potential of BEH and BH, we got some mixed liquors from BEH and BH respectively, then measured the redox potential by potentiometric titrator. The reaction between azo dyes and reduced electronic carrier is nonspecific reduction process. The occurrence of reaction was decided on the redox potential of redox intermediates and azo compound, namely determined by the redox potential (-180mV ~ -430mV) of cell redox cofactor, the NAD(P)H, and the potential of azo compound[9].TABLE II Rector BEH BH REDOX POTENTIAL Electric Current Density (mA/cm2) 0.024 0.048 0.071 Redox Potential (mV) -217 -185 -202 -152V. CONCLUSIONS • Compared with the single biological method, electrochemical function of DC micro-electric field and biological function produced synergy for the degradation processes of active brilliant red X-3B by bioelectrochemical technology. This method improved the activity of microorganism, and promoted the dye biodegradation. In the biodegradation process of azo dye active red X-3B by bioelectrochemical technology, the treatment effect increased with the increase of current density. The treatment efficiency of active brilliant red X-3B was increased with the increase of electric current density. The structure of active brilliant red X-3B was destroyed in bioelectrochemical hydrolysis reactor and turned into readily biodegradable small molecular organic compound. But in biological method, because the concentration of dye wastewater was too high, and the dissolved oxygen in hydrolysis reactor was lower, the bacteria in hydrolysis reactor were restrained. It made the structure of active brilliant red X-3B changed smaller.•••Table II was the redox potential of BEH and BH. Table II showed that, the redox potential of mixed liquor in BEH was about -200mV. It was in the range of azo compound redox requirement (-180mV ～ -430mV). The redox potential of mixed liquor in BH was about -152mV. It was not in the range of azo compound redox requirement, and also not in the range of hydrolysis reactor running normally (about 0mV). C. Azoreductase analysisThe first step of azo dyes biodegradation was the key to open azo double-bond, and produce aromatic ammonia. This step was finished by catalysis of azoreductase[10]. The decolorization ability of bacteria to the dyes mainly depended on the effect of azoreductase. This experiment analyzed azoreductase activity of the bacterium from each reactor. Measured each water sample in the same reactor three times, the average results were in table III.TABLE III BEH (mg/L·h) 55.33 AZOREDUCTASE ACTIVITY IN EACH REACTOR BCO （BEH-BCO） (mg/L·h) 3.17 BH (mg/L·h) 1.68 BCO (mg/L·h) 1.68As shown in table III, azoreductase activity in BEH-BCO system was higher than in BH-BCO system. The azoreductase activity in BEH reactor was 55.33 mg/L·h. It explained that due to the effect of direct current field, the rate of biochemical reactions was enhanced, microbial activity was better. It also explained that except for the electrochemical function and biodegradation function, the electric field could stimulate the activity of microbe and enhance the efficiency of dye degradation.Mi Yilei, Fan Jinhong, Ma Luming. Research on removal of azo dye by bioelectrochem ical technology [J]. Chinese Journal of Environmental Engineering, 2009, 3(8): 1457-1461. [2] Wang Hui. Recent Advance in Biological Treatment of Dyeing Wastewater [J]. Journal of Xia Men University (Natural Science), 2008, 12: 286-290. [3] Liang Hong, Zeng Kangmei. Progress in the Dyes Wastewater Treatment Processes [J]. Journal of Sichuan University of Science & Engineering, 2003, 6: 20-24. [4] Zhang Changsheng, Xue An, Zhao Huazhang. Progress in the studies of electrical bio-technology in environmental engineering[J]. Industrial Water Treatment, 2008, 28(3): 1-5. [5] Yao Guo, Wang Jian wei.Determination of Colority of Sewage Water[J]. PTCA (PART B: CHEM. ANAL.), 2008, 44:61-64. [6] Zhou Chunsheng,Yin Jun, Meng Lin. Study of Method for Determing TTC-Dehydrogenase [J].Journal of Jilin Architectural Civil Engineering Institute, 1995, 3 (1) [7] Brewster M, Fuss F, Tebbens J, et al. Spectrophotometer analysis of electrochemically treated simulated disperse dye bath effluent [M].AATCC, Book of papas, 1992: 17-19· [8] Zhao Yuhua ， Dong Ruijiao. Hydrolyzing Mechanism on Azo Dye Wastewater Treatment by Anaerobic Bioreactor [J]. Journal of Shenyang Jianzhu University (Natural Science), 2009, 25(2):325-328. [9] Liu Guangfei, Zhou Jiti, Wang Jing. Progress on Degradation of Azo Dyes by Bacteria and Azoreductase [J]. Environmental Science & Technology, 2006, 29(4): 112-114 [10] Ftrojt L， Strasak L， Vetted V， et a1． Comparison of the low—frequency magnetic field effects on bacteria Escherichia coli ， Leclercia adecarboxylata and Staphylococcus aureus[J] ．Bioelectrochemistry， 2004，63(12) ：337~341 [11] Chang Y H D， Grodzinsky A J ， Wang D I C ． Augmentation of masstransfer through electrical means for hydrogel-entrapped Escherichia colicultivation[J]．Biotechnol Bioeng， 1995， 48(2) ： 149~157 [12] Laleh Loghavi，B.S．electric field on growth kinetics，cellmembrane permeabilization，and frequencyresponse of microorganisms．The Ohio State University，2008[1]。
蓝梅主编 给排水科学与工程专业英语部分课文翻译中文版
第四单元给水系统一般来说，供水系统可划分为四个主要组成部分:（1）水源和取水工程（2）水处理和存储（3）输水干管和配水管网。
常见的未处理的水或者说是原水的来源是像河流、湖泊、泉水、人造水库之类的地表水源以及像岩洞和水井之类的地下水源。
修建取水构筑物和泵站是为了从这些水源中取水。
原水通过输水干管输送到自来水厂进行处理并且处理后的出水储存到清水池。
处理的程度取决于原水的水质和出水水质要求。
有时候，地下水的水质是如此的好以至于在供给给用户之前只需消毒即可。
由于自来水厂一般是根据平均日需求流量设计的，所以，清水池为水需求日变化量提供了一个缓冲区。
水通过输水干管长距离输送。
如果输水干管中的水流是通过泵所产生的压力水头维持的，那么我们称这个干管为增压管。
另外，如果输水干管中的水流是靠由于高差产生的可获得的重力势能维持的，那么我们称这个干管为重力管。
在输水干管中没有中间取水。
与输水干管类似，在配水管网中水流的维持要么靠泵增压，要么靠重力势能。
一般来说，在平坦地区，大的配水管网中的水压是靠泵提供的，然而，在不平坦的地区，配水管网中的压力水头是靠重力势能维持的。
一个配水管网通过引入管连接配水给用户。
这样的配水管网可能有不同的形状，并且这些形状取决于这个地区的布局。
一般地，配水管网有环状或枝状的管道结构，但是，根据当地城市道路和街区总体布局计划，有时候环状和枝状结构合用。
城市配水管网大多上是环状形式，然而，乡村地区的管网是枝状形式。
由于供水服务可靠性要求高，环状管网优于枝状管网。
配水管网的成本取决于对管网的几何形状合适的选择。
城市计划采用的街道布局的选择对提供一个最小成本的供水系统来说是重要的。
环状管网最常见的两个供水结构是方格状、环状和辐射状；然而，我们不可能找到一个最佳的几何形状而使得成本最低。
一般地，城镇供水系统是单入口环状管系统。
如上所说，环状系统有一些通过系统相互连接的管道使得通过这些连接接的管道，可以供水到同一个需水点。
给排水专业英语_英文1
Sewage treatmentSewage treatment, or domestic wastewatertreatment, is the process of removing contaminantsfrom wastewater and household sewage, bothrunoff (effluents) and domestic. It includesphysical, chemical, and biological processes toremove physical, chemical and biologicalcontaminants. Its objective is to produce a wastestream (or treated effluent) and a solid waste orsludge suitable for discharge or reuse back into theenvironment. This material is often inadvertentlycontaminated with many toxic organic andinorganic compounds.Origins of sewageSewage is created by residences, institutions, and commercial and industrial establishments. Raw influent (sewage) includes household waste liquid from toilets, baths, showers, kitchens, sinks, and so forth that is disposed of via sewers. In many areas, sewage also includes liquid waste from industry and commerce. The separation and draining of household waste into greywater and blackwater is becoming more common in the developed world, with greywater being permitted to be used for watering plants or recycled for flushing toilets. A lot of sewage also includes some surface water from roofs or hard-standing areas. Municipal wastewater therefore includes residential, commercial, and industrial liquid waste discharges, and may include stormwater runoff. Sewage systems capable of handling stormwater are known as combined systems or combined sewers. Such systems are usually avoided since they complicate and thereby reduce the efficiency of sewage treatment plants owing to their seasonality. The variability in flow also leads to often larger than necessary, and subsequently more expensive, treatment facilities. In addition, heavy storms that contribute more flows than the treatment plant can handle may overwhelm the sewage treatment system, causing a spill or overflow. It is preferable to have a separate storm drain system for stormwater in areas that are developed with sewer systems.As rainfall runs over the surface of roofs and the ground, it may pick up various contaminants including soil particles and other sediment, heavy metals, organic compounds, animal waste, and oil and grease. Some jurisdictions require stormwater to receive some level of treatment before being discharged directly into waterways. Examples of treatment processes used for stormwater include sedimentation basins, wetlands, buried concrete vaults with various kinds of filters, and vortex separators (to remove coarse solids).Process overviewSewage can be treated close to where it is created (in septic tanks, biofilters or aerobic treatment systems), or collected and transported via a network of pipes and pump stations to a municipal treatment plant (see sewerage and pipes and infrastructure). Sewage collection and treatment istypically subject to local, state and federal regulations and standards. Industrial sources of wastewater often require specialized treatment processes (see Industrial wastewater treatment).Conventional sewage treatment may involve three stages, called primary, secondary and tertiary treatment. Primary treatment consists of temporarily holding the sewage in a quiescent basin where heavy solids can settle to the bottom while oil, grease and lighter solids float to the surface. The settled and floating materials are removed and the remaining liquid may be discharged or subjected to secondary treatment. Secondary treatment removes dissolved and suspended biological matter. Secondary treatment is typically performed by indigenous, water-bornemicro-organisms in a managed habitat. Secondary treatment may require a separation process to remove the micro-organisms from the treated water prior to discharge or tertiary treatment. Tertiary treatment is sometimes defined as anything more than primary and secondary treatment. Treated water is sometimes disinfected chemically or physically (for example by lagoons and microfiltration) prior to discharge into a stream, river, bay, lagoon or wetland, or it can be used for the irrigation of a golf course, green way or park. If it is sufficiently clean, it can also be used for groundwater recharge or agricultural purposes.Process Flow Diagram for a typical large-scale treatment plantPre-treatmentPre-treatment removes materials that can be easily collected from the raw wastewater before they damage or clog the pumps and skimmers of primary treatment clarifiers (trash, tree limbs, leaves, etc).ScreeningThe influent sewage water is strained to remove all large objects carried in the sewage stream. This is most commonly done with an automated mechanically raked bar screen in modern plants serving large populations, whilst in smaller or less modern plants a manually cleaned screen may be used. The raking action of a mechanical bar screen is typically paced according to the accumulation on the bar screens and/or flow rate. The solids are collected and later disposed in a landfill or incinerated.Grit removalPre-treatment may include a sand or grit channel or chamber where the velocity of the incoming wastewater is carefully controlled to allow sand, grit and stones to settle.Primary treatmentIn the primary sedimentation stage,sewage flows through large tanks,commonly called "primary clarifiers" or"primary sedimentation tanks". The tanksare large enough that sludge can settle andfloating material such as grease and oilscan rise to the surface and be skimmed off.The main purpose of the primarysedimentation stage is to produce both agenerally homogeneous liquid capable ofbeing treated biologically and a sludgethat can be separately treated or processed.Primary settling tanks are usuallyequipped with mechanically drivenscrapers that continually drive the collected sludge towards a hopper in the base of the tank from where it can be pumped to further sludge treatment stages. Grease and oil from the floating material can sometimes be recovered for saponification.Secondary treatmentSecondary treatment is designed to substantially degrade the biological content of the sewage which are derived from human waste, food waste, soaps and detergent. The majority of municipal plants treat the settled sewage liquor using aerobic biological processes. For this to be effective, the biota require both oxygen and a substrate on which to live. There are a number of ways in which this is done. In all these methods, the bacteria and protozoa consume biodegradable soluble organic contaminants (e.g. sugars, fats, organic short-chain carbon molecules, etc.) and bind much of the less soluble fractions into floc. Secondary treatment systems are classified as∙fixed-film or∙suspended-growth.Fixed-film OR attached growth system treatment process including trickling filter and rotating biological contactors where the biomass grows on media and the sewage passes over its surface.In suspended-growth systems, such as activated sludge, the biomass is well mixed with the sewage and can be operated in a smaller space than fixed-film systems that treat the same amount of water. However, fixed-film systems are more able to cope with drastic changes in the amount of biological material and can provide higher removal rates for organic material and suspended solids than suspended growth systems.Roughing filters are intended to treat particularly strong or variable organic loads, typically industrial, to allow them to then be treated by conventional secondary treatment processes. Characteristics include typically tall, circular filters filled with open synthetic filter media to which wastewater is applied at a relatively high rate. They are designed to allow high hydraulic loading and a high flow-through of air. On larger installations, air is forced through the media using blowers. The resultant wastewater is usually within the normal range for conventional treatment processes.Activated sludgeMain article: Activated sludgeIn general, activated sludge plantsencompass a variety of mechanisms andprocesses that use dissolved oxygen topromote the growth of biological floc thatsubstantially removes organic material.The process traps particulate material andcan, under ideal conditions, convertammonia to nitrite and nitrate and ultimatelyto nitrogen gas, (see also denitrification).Surface-aerated basinsMost biological oxidation processesfor treating industrial wastewatershave in common the use of oxygen (orair) and microbial action.Surface-aerated basins achieve 80 to90% removal of Biochemical OxygenDemand with retention times of 1 to10 days. The basins may range indepth from 1.5 to 5.0 metres and usemotor-driven aerators floating on thesurface of the wastewater.In an aerated basin system, theaerators provide two functions: they transfer air into the basins required by the biological oxidation reactions, and they provide the mixing required for dispersing the air and for contacting the reactants (that is, oxygen, wastewater and microbes). Typically, the floating surface aerators are rated to deliver the amount of air equivalent to 1.8 to 2.7 kg O2/kW·h. However, they do not provide as good mixing as is normally achieved in activated sludge systems and therefore aerated basins do not achieve the same performance level as activated sludge units.Biological oxidation processes are sensitive to temperature and, between 0 °C and 40 °C, the rate of biological reactions increase with temperature. Most surface aerated vessels operate at between 4 °C and 32 °C.Filter beds (oxidizing beds)Main article: Trickling filterIn older plants and plants receiving more variable loads, trickling filter beds are used where the settled sewage liquor is spread onto the surface of a deep bed made up of coke (carbonized coal), limestone chips or specially fabricated plastic media. Such media must have high surface areas to support the biofilms that form. The liquor is distributed through perforated rotating arms radiating from a central pivot. The distributed liquor trickles through this bed and is collected in drains at the base. These drains also provide a source of air which percolates up through the bed, keeping it aerobic. Biologica l films of bacteria, protozoa and fungi form on the media’s surfaces and eat or otherwise reduce the organic content. This biofilm is grazed by insect larvae and worms which help maintain an optimal thickness. Overloading of beds increases the thickness of the film leading to clogging of the filter media and ponding on the surface.Biological aerated filtersBiological Aerated (or Anoxic) Filter (BAF) or Biofilters combine filtration with biological carbon reduction, nitrification or denitrification. BAF usually includes a reactor filled with a filter media. The media is either in suspension or supported by a gravel layer at the foot of the filter. The dual purpose of this media is to support highly active biomass that is attached to it and to filter suspended solids. Carbon reduction and ammonia conversion occurs in aerobic mode and sometime achieved in a single reactor while nitrate conversion occurs in anoxic mode. BAF is operated either in upflow or downflow configuration depending on design specified by manufacturer.Membrane bioreactorsMembrane bioreactors (MBR) combine activated sludge treatment with a membrane liquid-solid separation process. The membrane component uses low pressure microfiltration or ultra filtration membranes and eliminates the need for clarification and tertiary filtration. The membranes are typically immersed in the aeration tank; however, some applications utilize a separate membrane tank. One of the key benefits of an MBR system is that it effectively overcomes the limitations associated with poor settling of sludge in conventional activated sludge (CAS) processes. The technology permits bioreactor operation with considerably higher mixed liquor suspended solids (MLSS) concentration than CAS systems, which are limited by sludge settling. The process is typically operated at MLSS in the range of 8,000–12,000 mg/L, while CAS are operated in the range of 2,000–3,000 mg/L. The elevated biomass concentration in the MBR process allows for very effective removal of both soluble and particulate biodegradable materials at higher loading rates. Thus increased Sludge Retention Times (SRTs) — usually exceeding 15 days — ensure complete nitrification even in extremely cold weather.The cost of building and operating an MBR is usually higher than conventional wastewater treatment. Membrane filters can be blinded with grease or abraded by suspended grit and lack a clarifier's flexibility to pass peak flows. The technology has become increasingly popular for reliably pretreated waste streams and has gainedwider acceptance where infiltration and inflowhave been controlled, however, and the life-cyclecosts have been steadily decreasing. The smallfootprint of MBR systems, and the high qualityeffluent produced, make them particularly usefulfor water reuse applications.There are MBR plants being built throughout theworld, including North Librty, Iowa, Georgia, andCanada.Secondary sedimentationThe final step in the secondary treatment stage is to settle out the biological floc or filter material and produce sewage water containing very low levels of organic material and suspended matter.Rotating biological contactorsMain article: Rotating biological contactorRotating biological contactors(RBCs) are mechanical secondarytreatment systems, which arerobust and capable ofwithstanding surges in organicload. RBCs were first installed inGermany in 1960 and have sincebeen developed and refined into areliable operating unit. Therotating disks support the growthof bacteria and micro-organismspresent in the sewage, whichbreakdown and stabilise organicpollutants. To be successful, micro-organisms need both oxygen to live and food to grow. Oxygen is obtained from the atmosphere as the disks rotate. As the micro-organisms grow, they build up on the media until they are sloughed off due to shear forces provided by the rotating discs in the sewage. Effluent from the RBC is then passed through final clarifiers where the micro-organisms in suspension settle as a sludge. The sludge is withdrawn from the clarifier for further treatment.A functionally similar biological filtering system has become popular as part of home aquarium filtration and purification. The aquarium water is drawn up out of the tank and then cascaded over a freely spinning corrugated fiber-mesh wheel before passing through a media filter and back into the aquarium. The spinning mesh wheel develops a biofilm coating of microorganisms that feed on the suspended wastes in the aquarium water and are also exposed to the atmosphere as the wheel rotates. This is especially good at removing waste urea and ammonia urinated into the aquarium water by the fish and other animals.。
给排水专业英语翻译
Particle-ParticleInteractionsParticle-particle interactions are extremely important in bringing about aggregation by means of Brownian motion .the theory that has been developed to describe particle-particle interactions is based on the consideration of interaction between two charged flat plates and between two charged spheres.As neither of these developments is directly applicable to the particles found in wastewater ,as described above,the analysis for two charged flat plates will be used for illustrative purposes.The two principal face involved are the forces of repulsion,due to the electrical properties of the charged piates,and the van der Waals forces of attraction.It should be noted that the van der Waals forces of attraction do not come into play until the two plates are brought together in close proximity to each other.The net total energy shown by the solid lines on Figure 3 is the difference between the forces of repulsion and attraction.The two conditions,with respect to the forces of repulsion ,are illustrated onFigure 3.as shown for conditions 1, the forces ofattraction will predominate at short and long distances.The net energy curve for condition 1containsa repulsive maximum that must be overcome if theparticles,represented as the two plates,are to be heldtogether by the van der Waals force of attraction.Incondition 2,there is no energy barrier toovercome .Clearly ,if colloidal particles are to beremoved by microflocculation,the repulsive force mustbe reduced.although floc particles can form at alongdistances as shown by the energy curve for condition1, the net force holding these particles together isweak and the floc particles that are formed can beruptured easily. Particle Destabilization with Potential-Determining Ions and ElectrolytesTo bring about particle aggregation through microflocculation,steps must be taken to reduceparticle charge or to overcome the effect of thischarge.the effect of the charge can be overcome by(1)the addition of potential-determining ions,which will be taken up by or will react with the colloid surface to lessen the surfaceto lessen chargeand(2)the addition f electrolytes,which have theeffect of reducing thethickness of the diffuse electric layer and, thereby,reduce the zeta potentialUse of Potential-Determining Ions.The addition of Potential-Determining ions to promotecoagulaion can be illustrated by theaddition of strong acids or bases to reduethe charge of metal oxides or hydroxides to near zero so that coagulation can occur.The effect of addingpotential-determining ions in a solution containing charged particles is illustrated on figure4.The magnitude of the effect will depend on the conc entration of potential-determining ionsadded.The f ollowing ratios,known as the Shultz-Hardy rule,can be used to assess theeffectiveness of potential-determining or counterions:1:1/2^6:1/3^6or100;1.6:0.13(3)It is interesting to note that depending on the concentration and nature of the counterions added,it is possible to reverse the charge of the double layer a nd develop a new stable particle.The effect of adding counterions to a solution containing charged particles is illustrated on Figure 5.The upper curve on Figure 5 represents the surface charge of the particle as a function of the concentration of counterions added .The lines designated kT represent the thermal kinetic energy of the particle .The lower diagram is a plot of the turbidity that would result if the particles that have been destabilized and have undergone microflocculation were removed by settling .As shown ,when the surface charge (either positive or negative)is greater than the thermal kinetic energy of the particles ,the particles will not flocculate and original turbidity is observed.Additional detail on the use of counterions may be found in Shaw(1966).The use of potential determining ions is not feasible in either water or wastewater treatment because of the massive concentration of ions that must be added to bring about sufficient compression of the electrical double layer to effect perikinetic flocculatin.Use of Electrolytes .Electrolytes can also be added to coagulate colloidal suspensions .Increased concentration of a given electrolyte will cause a decrease in zeta potential and a corresponding decreasein repulsive forces as illustrated in condition 2 on Figure 3 and on Figure 4.The concentration of an electrolyte that is needed to destabilize a colloidal suspension is known as the critical coagulation concentration(ccc).Increasing the concentration of an indifferent electrolyte will not result in the restabitization of the colloidal particles .As with the addition of potential-determining ions ,the use of electrolytes is also not feasible in wastewater treatment .As discussed subsequently ,a change in the particle charge will occur when chemicals are added to adjust the pH of the wastewater to optimize the performance of hydrolyzed metal ions used as coagulants.Particle Destabilization and Aggregation with Polyelectrolytes Polyelectrolytes may be divided into two categories :natural and synthetic .Important naturalpolyelectrolytes include polymers of biological origin and those derived for starch products such as cellulose derivatives and alginates .Synthetic polyelectrolytes consist of simple monomers that are polymerized into high-molecular-weight substances .Depending on whether their charge ,when placed in water ,is negative .positive ,or neutral ,these polyelectrolytes are classified as anionic ,cationic , and nonionic ,respectively .The action of polyelectrolytes may be divided into the following three general categories.Charge Neutralization .In the first category ,polyelectrolytes act as coagulants that neutralize or lower the charge of the wastewater particles .Because wastewater particle s normally are charged negatively ,cationic polyelectrolytes are used for this purpose . In this application ,the cationic polyelectrolytes are considered to be primary coagulants .To effect charge neutralization ,the polyelectrolyte must be adsorbed to the particle .Because of the large number of particles found in wastewater ,the mixing intensity must besufficient to bring about the adsorption of the polymer onto the colloidal particles .With inadequate mixing ,the polymer will eventually fold back on itself and its effectiveness in reducing the surface charge will be diminished .Further , if the number of colloidal particles is limited ,it will be difficult to remove them with low polyelectrolyte dosages .Polymer Bridge Formation . The second mode of action of polyelectrolytes is interparticle bridging (see Figure 6) .In this case ,polymers that are anionic (usually anionic to a slight extent when placed in water) become attached at a number of adsorption sites to the surface to the surface of the particles found in the wastewater .A bridge is formed when two or more particles become adsorbed along the length of the polymer .。
给排水外文翻译
给排水外文翻译【概述】外文名称：Water Supply and Drainage【引言】给排水是指人类为了满足生活、生产和环境需求，采集、利用和排放水资源的活动和系统。
随着城市化进程的加速和人们对舒适生活品质的要求不断提高，给排水工程在城市规划和建设中起到至关重要的作用。
本文将介绍给排水外文翻译的重要性、翻译技巧和注意事项，为给排水工程相关专业人员提供参考。
【翻译重要性】给排水工程涉及大量外文文献和技术资料，而国内外水利工程界的发展迅猛，相关外文文献的翻译对于我国的给排水工程建设具有重要意义。
通过翻译，我们可以了解国外先进的给排水技术和管理经验，为我国的工程建设提供参考和借鉴。
同时，翻译还有助于加强国际间的交流合作，促进我国在给排水领域的影响力和地位的进一步提升。
【翻译技巧】1. 理解专业术语：给排水领域涉及大量专业术语，翻译者应对这些术语进行准确理解。
可以通过查阅外文词典或专业词汇表对其进行翻译，避免出现术语误译的情况。
2. 深入研究上下文：在翻译过程中，翻译者应该深入研究原文上下文，理解全文的语境和主旨，以确保翻译结果的准确性和一致性。
3.注意句子结构：外文论文的句子结构和汉语差异较大，翻译者应根据汉语表达习惯进行适当调整，保证译文通顺。
【注意事项】1. 外文翻译要准确传达论文内容，不得随意增删原文内容。
2. 翻译过程中应注意句子结构的转换，确保译文的准确性和流畅性。
3. 注意专业术语翻译的准确性，可以参考国内外相关词汇表和标准进行翻译。
4. 翻译过程中应注意时间和质量的把握，提前制定翻译计划，并进行分段、分步翻译，以确保高质量的翻译成果。
【结论】给排水外文翻译对于我国给排水工程建设和国际交流具有重要意义。
翻译者需要具备扎实的专业知识和翻译技巧，通过深入研究与准确翻译，为我国的工程建设和国际交流贡献力量。
同时，加强对外文文献和技术的翻译工作，不断提高我国在给排水领域的创新能力和核心竞争力，助力我国以科技创新引领未来社会发展的目标实现。