Wastewater Engineering: Advanced Wastewater Treatment Systems

Wastewater Engineering: Advanced Wastewater Treatment Systems

© IJSR Publications

Wastewater Engineering: Advanced Wastewater Treatment Systems

ISSN: 2322-4657 DOI 10.12983/1-2014-03-01 © IJSR Publications, Penang, Malaysia,

This work is subject to copyright. All rights are reserved by the Publisher, whether the whole or part of the material is concerned, specifically the rights of translation, reprinting, reuse of illustrations, recitation, broadcasting, reproduction on microfilms or in any other physical way, and transmission or information storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology now known or hereafter developed. Exempted from this legal reservation are brief excerpts in connection with reviews or scholarly analysis or material supplied specifically for the purpose of being entered and executed on a computer system, for exclusive use by the purchaser of the work. Duplication of this publication or parts thereof is permitted only under the provisions of the Copyright Law of the Publisher’s location, in its current version, and permission for use must always be obtained from IJSR Publications. Permissions for use may be obtained through RightsLink at the Copyright Clearance Center. Violations are liable to prosecution under the respective Copyright Law. The use of general descriptive names, registered names, trademarks, service marks, etc. in this publication does not imply, even in the absence of a specific statement, that such names are exempt from the relevant protective laws and regulations and therefore free for general use. While the advice and information in this book are believed to be true and accurate at the date of publication, neither the authors nor the editors nor the publisher can accept any legal responsibility for any errors or omissions that may be made. The publisher makes no warranty, express or implied, with respect to the material contained herein.

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ii

PREFACE

As the global population grows and many developing countries modernize, the importance of water supply and wastewater treatment becomes a much greater factor in the welfare of nations. Clearly, in today’s world the competition for water resources coupled with the unfortunate commingling of wastewater discharges with freshwater supplies creates additional pressure on treatment systems. Recently, researchers focus on wastewater treatment by difference methods with minimal cost and maximum efficiency. This volume of the Wastewater Engineering: Advanced Wastewater Treatment Systems is a selection of topics related to physical-chemical and biological processes with an emphasis on their industrial applications. It gives an overview of various aspects in wastewater treatments methods including topics such as biological, bioremediation, electrochemical, membrane and physical-chemical applications. Experts in the area of environmental sciences from diverse institutions worldwide have contributed to this book, which should prove to be useful to students, teachers, and researchers in the disciplines of wastewater engineering, chemical engineering, environmental engineering, and biotechnology. We gratefully acknowledge the cooperation and support of all the contributing authors.

Hamidi Abdul Aziz

Amin Mojiri

Professor, School of Civil Engineering,

Research Assistant, School of Civil Engineering,

Engineering Campus, Universiti Sains Malaysia,

Engineering Campus, Universiti Sains Malaysia,

[email protected]

[email protected]

iii

TABLE OF CONTENTS

PREFACE ............................................................................................................................... iii TABLE OF CONTENTS ........................................................................................................ iv CHAPTER

1:

INTRODUCTION

OF

PRELIMINARY

AND

SECONDARY

TREATMENTS ...................................................................................................................... 1 1.1 Introduction of preliminary and Secondary Treatments; Z. Amirossadat .............. 2 CHAPTER 2: WASTEWATER BY TREATMENT BY PHYSICAL-CHEMICAL TECHNOLOGIES ............................................................................................................... 05 2.1 Recent Development in Landfill Leachate Treatment Using Low Cost Adsorbent Prepared From Waste Material; N. Azmi, J.K. Bashir, S. Sethupathi, C.A. Ng ......... 06 2.2 Removal of Colour from Synthetic Dye Wastewater Using Adsorbent Prepared from Psyllium Husk; I. Dahlana and S.M.O. Tayeh ............................................................. 15 2.3 COD and BOD Removal from Textile Wastewater Using Naturally Prepared Adsorbents and Their Activation forms Using Sulphuric Acid; Patel and Vashi ........ 31 2.4 Fenton oxidation for the Treatment of Liquid Waste with High COD and Anionic/Non-ionic Surfactants; M. Collivignarelli, S. Sorlini, A. Abbà, M. Sordi ..... 41 2.5 Ultrasound Irradiation on the Treatment of Aromatic Compounds in Wastewater W.L. Peng, G. Xinxin, M.J.K. Bashir .......................................................................... 48 CHAPTER 3: WASTEWATER TREATMENT BY BIOLOGICAL METHODS ........ 62 3.1 Wastewater Treatment by Biological Methods; A. Dadrasnia, N. Shahsavari and C.U. Emenike ............................................................................................................... 63 3.2 Biological Treatment of Recycled Paper Mill Wastewater Using Modified Anaerobic Inclining-Baffled (MAIB) Bioreactor; H.M. Zwaina and I. Dahlan ............................ 71 3.3 Augmentation of Biological Nitrogen Removal via Optimization of Support Media Size and Aeration Strategy in Moving Bed Sequencing Batch Reactor; J.Wei Lim, M.J.K. Bashir, S.L. Ng, S. Sethupathi, L.P. Wong. ................................................................. 87

iv

CHAPTER 4: ELECTROCHEMICAL METHODS ........................................................ 96 4.1 Electrochemical Oxidation Process Contribution in Remediating Complicated Wastewaters; M.J. K. Bashir, J.W. Lim, S.Q. Aziz, S.S.A. Amr ................................ 97 CHAPTER

5:

WASTEWATER

TREATMENT

BY

BIOREMEDIATION

TECHNOLOGIES ............................................................................................................. 107 5.1 Wastewater Treatment by Bioremediation Methods; A.N. Amenaghawon and K.O. Obahiagbon ................................................................................................................ 108 5.2 Supplementation of Novel Solid Carbon Source Prepared from Dried AttachedGrowth Biomass for Bioremediation of Wastewater Containing Nitrogen; J.W. Lim, M.J.K. Bashir, C.A. Ng, X. Guo. ............................................................................... 125 CHAPTER 6: WASTEWATER TREATMENT BY MEMBRANE TECHNIQUES . 136 6.1 Supported Liquid Membrane in wastewater Treatment; T.T. Teng, A. Talebi, and G. Muthuraman ............................................................................................................... 137 6.2 Role of Emulsion Liquid Membrane (ELM) in Separation Processes; T.T. Teng, M. Soniya, G. Muthuraman and A. Talebi ...................................................................... 149 6.3 Bulk Liquid Membrane and its Applications in Wastewater Treatment; T.T. Teng, S. Elumalai, G. Muthuraman and A. Talebi ................................................................... 158 6.4 Challenges in Fabricating Suitable Membrane for Water Treatment Application; L.Y. Wong, C.A. Ng, MJ.K. Bashir, T.L. Chew ....................................................... 171 6.5 Removal of Copper from Aqueous Solution by Using Thermo-responsive Polymer Hydrogel as Copper Chelator via Polymer-enhanced Ultrafiltration (PEUF); J.J. Chen, A.L. Ahmad and B.S. Ooi .......................................................................................... 183 CHAPTER

7:

WASTEWATER

TREATMENT

BY

PHYTOREMEDIATION

TECHNOLOGIES ............................................................................................................. 193 7.1 Wastewater Treatment by Phytoremediation Methods; H. Farraji..................... 194 CHAPTER 8: LANDFILL LECHATE TREATMENT TECHNIQUES...................... 207 8.1 Municipal Landfill Leachate Treatment Techniques: An Overview; S.Q. Aziz, H.A. Aziz, M.J.K. Bashir, A. Mojiri .................................................................................. 208

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CHAPTER 9: APPLICATION OF OPTIMIZATION IN TREATMENT................... 225 9.1 Application of Optimization in Wastewater Treatment; Y.L. Lim, Y.C. Ho, A.F.M. Alkarkhi ..................................................................................................................... 226

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Wastewater Engineering: Advanced Wastewater Treatment Systems

Chapter 1: Introduction of preliminary and Secondary Treatments

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Introduction of preliminary and Secondary Treatment Zahra Amirossadat Isfahan (Khorasgan) Branch, Islamic Azad University, Isfahan, Iran Abstract. Recently, the amounts of wastewater are sharply increasing and the kinds of pollutants are also varied as the world wide industry is being developed incessantly. With respect to both the quantity and composition, the textile processing wastewater is recorded as the most polluted source among all industrial sectors. This chapter explained the preliminary and secondary treatment of wastewater. Keywords: Preliminary treatment, Secondary treatment, Wastewater

1. INTRODUCTION Recently, the amounts of wastewater are sharply increasing and the kinds of pollutants are also varied as the world wide industry is being developed incessantly. With respect to both the quantity and composition, the textile processing wastewater is recorded as the most polluted source among all industrial sectors (Chang et al., 2009). At wastewater treatment plants, wastewater is treated before it is allowed to be returned to the environment, lakes, or streams. Discharge criteria required the installation of facilities that performed what is now called primary treatment of wastewater. This involved using screens and sedimentation tanks to remove most of the materials in the wastewater that float or settle. As subsequent discharge criteria were tightened, secondary treatment became necessary. Secondary treatment is accomplished by bringing together waste, bacteria and oxygen in trickling filters or the activated sludge process. Bacteria are used to consume the organic parts of the wastewater. Facilities and their designers are now considering and installing tertiary treatment facilities to comply with the latest regulatory and permit parameters. These advanced treatment processes go beyond conventional secondary treatment and include the removal of recalcitrant organic compounds, as well as excess nutrients such as nitrogen and phosphorus (Coppen, 2004). Conventional wastewater treatment consists of a combination of physical, chemical, and biological processes and operations to remove solids, organic matter and, sometimes, nutrients from wastewater. General terms used to describe different degrees of treatment, in order of increasing treatment level, are preliminary, primary, secondary, and tertiary and/or advanced wastewater treatment. In some countries, disinfection to remove pathogens sometimes follows the last treatment step.

2. Primary treatment The objective of primary treatment is the removal of settleable organic and inorganic solids by sedimentation, and the removal of materials that will float (scum) by skimming. Approximately 25 to 50% of the incoming biochemical oxygen demand (BOD5), 50 to 70% of the total suspended solids (SS), and 65% of the oil and grease are removed during primary treatment. Some organic nitrogen, organic phosphorus, and heavy metals associated with solids are also removed during primary sedimentation but colloidal and dissolved constituents are not affected. The effluent from primary sedimentation units is referred to as primary effluent. In many industrialized countries, primary treatment is the minimum level of reapplication treatment required for wastewater irrigation. It may be considered sufficient treatment if the wastewater is used to irrigate crops that are not consumed by humans or to irrigate orchards, vineyards, and some processed food crops. However, to prevent potential nuisance conditions in storage or flow-equalizing reservoirs, some form of secondary treatment is normally required in these countries, even in the case of non-food crop irrigation. It may be possible to use at least a portion of primary effluent for irrigation if off-line storage is provided. Primary sedimentation tanks or clarifiers may be round or rectangular basins, typically 3 to 5 m deep, with hydraulic retention time between 2 and 3 hours. Settled solids (primary sludge) are normally removed from the bottom of tanks by sludge rakes that scrape the sludge to a central well from which it is pumped to sludge processing units. Scum is swept across the tank surface by water jets or mechanical means from which it is also pumped to sludge processing units. (http://www.fao.org/docrep/t0551e/t0551e05.htm). Primary treatment involves:

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Amirossadat Advanced Wastewater Treatment

1. Screening- to remove large objects, such as stones or sticks that could plug lines or block tank inlets. 2. Grit chamber- slows down the flow to allow grit to fall out 3. Sedimentation tank (settling tank or clarifier)settleable solids settle out and are pumped away, while oils float to the top and are skimmed off (http://www.sd1.org/resourcehandler.aspx?id=28).

or ultraviolet light). An increasing number of wastewater facilities also employ tertiary treatment, often using advanced treatment methods. Tertiary treatment may include processes to remove nutrients such as nitrogen and phosphorus, and carbon adsorption to remove chemicals. These processes can be physical, biological, or chemical. Settled solids (sludge) from primary treatment and secondary treatment settling tanks are given further treatment and undergo several options for disposal (http://www.sd1.org/resourcehandler.aspx?id=28). The objective of secondary treatment is the further treatment of the effluent from primary treatment to remove the residual organics and suspended solids. In most cases, secondary treatment follows primary treatment and involves the removal of biodegradable dissolved and colloidal organic matter using aerobic biological treatment processes. Aerobic biological treatment (see Box) is performed in the presence of oxygen by aerobic microorganisms (principally bacteria) that metabolize the organic matter in the wastewater, thereby producing more microorganisms and inorganic end-products (principally CO2, NH3, and H2O). Several aerobic biological processes are used for secondary treatment differing primarily in the manner in which oxygen is supplied to the microorganisms and in the rate at which organisms metabolize the organic matter. High-rate biological processes are characterized by relatively small reactor volumes and high concentrations of microorganisms compared with low rate processes. Consequently, the growth rate of new organisms is much greater in high-rate systems because of the well-controlled environment. The microorganisms must be separated from the treated wastewater by sedimentation to produce clarified secondary effluent. The sedimentation tanks used in secondary treatment, often referred to as secondary clarifiers, operate in the same basic manner as the primary clarifiers described previously. The biological solids removed during secondary sedimentation, called secondary or biological sludge, are normally combined with primary sludge for sludge processing. Common high-rate processes include the activated sludge processes, trickling filters or biofilters, oxidation ditches, and rotating biological contactors (RBC). A combination of two of these processes in series (e.g., biofilter followed by activated sludge) is sometimes used to treat municipal wastewater containing a high concentration of organic material from industrial sources.

2. Secondary treatment Secondary wastewater treatment is the second stage of wastewater treatment that takes place after the primary treatment process. The process consists of removing or reducing contaminants or growths that are left in the wastewater from the primary treatment process. Usually biological treatment is used to treat wastewater in this step because it is the most effective type of treatment on bacteria, or contaminant, growth. Secondary treatment processes can remove up to 90 percent of the organic matter in wastewater by using biological treatment processes. The two most common conventional methods used to achieve secondary treatment are attached growth processes and suspended growth processes (http://www.water.siemens.com/en/applications/waste water_treatment/secondary-treatment). Secondary treatment typically utilizes biological treatment processes, in which microorganisms convert nonsettleable solids to settleable solids. Sedimentation typically follows, allowing the settleable solids to settle out. Three options include: 1. Activated Sludge- The most common option uses microorganisms in the treatment process to break down organic material with aeration and agitation, then allows solids to settle out. Bacteria-containing “activated sludge” is continually recirculated back to the aeration basin to increase the rate of organic decomposition. 2. Trickling Filters- These are beds of coarse media (often stones or plastic) 3-10 ft. deep. Wastewater is sprayed into the air (aeration), then allowed to trickle through the media. Microorganisms attached to and growing on the media, break down organic material in the wastewater. Trickling filters drain at the bottom; the wastewater is collected and then undergoes sedimentation. 3. Lagoons- These are slow, cheap, and relatively inefficient, but can be used for various types of wastewater. They rely on the interaction of sunlight, algae, microorganisms, and oxygen (sometimes aerated). After primary and secondary treatment, municipal wastewater is usually disinfected using chlorine (or other disinfecting compounds, or occasionally ozone

(a) Activated Sludge In the activated sludge process, the dispersed-growth reactor is an aeration tank or basin containing a

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Wastewater Engineering: Advanced Wastewater Treatment Systems Chapter 1: Introduction of preliminary and Secondary Treatments

suspension of the wastewater and microorganisms, the mixed liquor. The contents of the aeration tank are mixed vigorously by aeration devices which also supply oxygen to the biological suspension. Aeration devices commonly used include submerged diffusers that release compressed air and mechanical surface aerators that introduce air by agitating the liquid surface. Hydraulic retention time in the aeration tanks usually ranges from 3 to 8 hours but can be higher with high BOD5 wastewaters. Following the aeration step, the microorganisms are separated from the liquid by sedimentation and the clarified liquid is secondary effluent. A portion of the biological sludge is recycled to the aeration basin to maintain a high mixed-liquor suspended solids (MLSS) level. The remainder is removed from the process and sent to sludge processing to maintain a relatively constant concentration of microorganisms in the system. Several variations of the basic activated sludge process, such as extended aeration and oxidation ditches, are in common use, but the principles are similar.

increases as new organisms grow. Periodically, portions of the film 'slough off the media. The sloughed material is separated from the liquid in a secondary clarifier and discharged to sludge processing. Clarified liquid from the secondary clarifier is the secondary effluent and a portion is often recycled to the biofilter to improve hydraulic distribution of the wastewater over the filter. (c) Rotating Biological Contactors Rotating biological contactors (RBCs) are fixed-film reactors similar to biofilters in that organisms are attached to support media. In the case of the RBC, the support media are slowly rotating discs that are partially submerged in flowing wastewater in the reactor. Oxygen is supplied to the attached biofilm from the air when the film is out of the water and from the liquid when submerged, since oxygen is transferred to the wastewater by surface turbulence created by the discs' rotation. Sloughed pieces of biofilm are removed in the same manner described for biofilters (http://www.fao.org/docrep/t0551e/t0551e05.htm).

(b) Trickling Filters A trickling filter or biofilter consists of a basin or tower filled with support media such as stones, plastic shapes, or wooden slats. Wastewater is applied intermittently, or sometimes continuously, over the media. Microorganisms become attached to the media and form a biological layer or fixed film. Organic matter in the wastewater diffuses into the film, where it is metabolized. Oxygen is normally supplied to the film by the natural flow of air either up or down through the media, depending on the relative temperatures of the wastewater and ambient air. Forced air can also be supplied by blowers but this is rarely necessary. The thickness of the biofilm

REFERENCES Chang W, Tran H, Park D, Zhang R, Ahn D (2009). Ammonium nitrogen removal characteristics of zeolite media in a Biological Aerated Filter (BAF) for the treatment of textile wastewater. Journal of Industrial and Engineering Chemistry, 15: 524-528. Cppen J (2004) Advanced Wastewater Treatment Systems. Courses ENG4111 and 4112 Research Project, University of Southern Queensland, Faculty of Engineering and Surveyin.

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Wastewater Engineering: Advanced Wastewater Treatment Systems

Chapter 2: Wastewater Treatment by PhysicalChemical Technologies

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Recent Development in Landfill Leachate Treatment Using Low Cost Adsorbent Prepared From Waste Material Nurshazwani Binti Azmi, Mohammed J.K. Bashir*, Sumathi Sethupathi, Choon-Aun Ng Department of Environmental Engineering, Faculty of Engineering and Green Technology (FEGT), University Tunku Abdul Rahman, 31900 Kampar, Perak, Malaysia *Corresponding Author: [email protected]; Tel: 605-4688888 ext: 4559; Fax: 605-4667449 Abstract. Landfill leachate has become the subject of recent research interest as it is a strongly polluted wastewater. The produced leachate is one of the most important drawbacks of municipal solid waste disposal (MSW) in sanitary landfill. Adsorption by activated carbon (AC) appears to have considerable potential in landfill leachate treatment due to the simplicity design, superior removal of organic compound and less land area required. However, the high demand for AC is a major problem due to limited carbon based substances such as coal, wood and lignite. Therefore, waste material seems to be a good option as an alternatives source of AC. Consequently, this paper focuses on effectiveness of using AC in landfill leachate treatment and highlighted the recent development treating landfill leachate using adsorbent prepared from waste material. Keywords: Landfill Leachate, Treatment, Adsorbent

inorganic salts (Renou et al., 2008). Without an appropriate treatment, landfill leachate could be a potential source of surface and groundwater contamination, as it could seep into soils and subsoil, causing severe pollutions to receiving water body (Oman and Junestedt, 2008). Typically, leachate characteristics and compositions depends on various factors such as waste composition, age of landfill, site hydrogeology, specific climate conditions, moisture routing through the landfill, and the landfill design and operation (Ghafari et al., 2010). Age of landfill site is one of the main variables that affect the leachate characteristics (Bashir et al., 2012), where the concentration of leachates parameters changes with the age of the leachate. Young acidogenic landfill leachate commonly characterized by high biochemical oxygen demands(BOD) and chemical oxygen demands(COD) , high concentration of ammonium nitrogen followed by low pH value as low as pH 4 (Wu et al., 2001). The degradation of biological matter by microorganism lead to generation of Volatile Fatty Acid(VFA) that lead to low ph value and high BOD/COD ratio. On the contrary, aged landfill (i.e. >10 years old) produces mature (stabilized) leachate that contains bio-refractory compounds such as humic acid (HA) and fulvic acids (FA), with BOD5/COD ratio less than 0.1 (Alvariz-Vazqurez et al., 2004) as illustrated in Table 1. Biological treatment of landfill leachates have been shown to be very effective in removing organic matter in early stages (Berruetta and Castrillon, 1992) with high BOD/COD ratio. As the BOD/COD ratio decrease with the passage of time (Rodriguez et al., 2000), the biodegradable organic content of leachate reduced where biological treatment no longer effective due to the presence of refractory organic matter and physico-chemical processes may become

1. INTRODUCTION As the exponential population and social civilization growth, together with the developments of industries and technologies, rapid generation of MSW has becomes a global environmental problem (Saeed et al., 2009). There are many options available for MSW disposal such as sanitary landfill, open dump, incineration, composting, grinding, hog feeding, milling, and anaerobic digestion (Aziz et al., 2010). Sanitary landfill is the most common MSW disposal method due to such advantages as the simple disposal procedure, low cost, and landscape-restoring effect on holes from mineral workings (Bashir et al., 2010). However, the production of highly contaminated leachate is a major drawback of this method (Aziz et al., 2010). Landfill leachate is defined as any contaminated liquid effluent percolating through deposited waste and emitted within a landfill or dump site through external sources (Taulis, 2005). In a more precise definition, it is a soluble organic and mineral compound formed when water infiltrates into the refuse layers, which extracts a series of contaminants and instigates a complex interplay between the hydrological and biogeochemical reactions that acts as a mass transfer mechanisms for producing of moisture content sufficiently high to start the liquid flow (Aziz et al., 2004). As shown in Figure 1, leachate generation induced by the gravity force, precipitation, irrigation, surface runoff, rainfall, snowmelt, recirculation, liquid waste co-disposal, refuse decomposition, groundwater intrusion and initial moisture content present within the landfills (Achankeng, 2004). As the consequences, leachate may contain high concentration of organic matter (biodegradable and non-biodegradable), ammonia nitrogen, heavy metals, chlorinated organic and

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Azmi et al. Recent Development in Landfill Leachate Treatment Using Low Cost Adsorbent Prepared From Waste Material

one of the appropriate options for stabilized landfill leachate. Various physico-chemical process have been practiced for old landfill leachate treatments such as adsorption (Halim et al., 2012), ion-exchange (Bashir et al., 2012), Fenton reaction (Mohajeri et al., 2011), coagulation/flocculation (Ghafari et al., 2010), electrochemical oxidation process (Marco et al., 2013) ozonation (Salem et al., 2013), and air stripping (Bloor and Banks, 2005). Among all process, adsorption technology is one of the most applicable and simple methods. Adsorption is defined as a mass transfer process by which a substance is transferred from liquid or gas phases to the solid surface of adsorbent and form attachment via physical or chemical interactions. The material providing the solid surface is called the adsorbent and material removed from the liquid phase is called as adsorbate. AC demonstrated significant adsorption efficiency in gas and liquid phases due to its high micropore volume, large specific surface area,

favorable pore size distribution, thermal stability, and capability for rapid adsorption and low acid/base reactivity (Li et al., 2009). The unique adsorptive properties of AC, makes it as one of the best filtration media in the world. However, high manufacturing cost and expensive carbonaceous material for producing high quality AC (Mohan and Pittman, 2006) lead to limitation of this application for landfill leachate treatment especially in developing countries. Thus, the use of non-conventional material such as agriculture waste and industrial by-product that are locally available can be chemically modified and utilized as a low carbon adsorbent (Babel and Kurniawan, 2003). Several studies have been conducted by using AC for various types of waste water. Consequently, the present work reviews and evaluates the recent published works focuses on landfill leachate treatment using adsorbent prepared from the waste materials.

Fig. 1: Leachate Formation (Agamuthu, 2001)

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Wastewater Engineering: Advanced Wastewater Treatment Systems Chapter 2: Wastewater Treatment by Physical-Chemical Technologies

Table 1: Classification of landfill leachate (Alvarez-Vazquez et al., 2004) Parameters Age (years)

Young 10

pH

7.5

COD (mg L-1)

>10 000

4000-10 000

chromium. They further reported positive change in entropy for each metal and the order of disorder was nickel>cobalt >chromium. Lodeiro et al. (2005) investigated the potential use of five different brown seaweeds, Bifurcaria bifurcata, Saccorhiza polyschides, Ascophyllum nodosum, Laminaria ochroleuca and Pelvetia

caniculata for the removal of cadmium from aqueous solutions. They observed that the biosorption process was relatively fast with about 90% removal of cadmium occurring within 1 hour. Chen and Wang, (2008) investigated the removal of lead, silver, caesium and strontium from aqueous solution using brewery's waste biomass. Their results revealed that the biosorption process was rapid and was well described by the pseudo second order kinetic and Langmuir isotherm models. Ho, (2005) investigated the biosorption of lead using tree fern in a baffled agitated system. The optimum pH for lead removal was determined to be 4.9. The pseudo second order kinetic model sufficiently described the kinetics of the biosorption process. Bishnoi et al. (2004) studied the removal of chromium from aqueous solutions using activated rice husk and activated alumina while Garg et al. (2004) studied the removal of chromium from aqueous solution using formaldehyde treated saw dust and saw dust carbon activated with sulphuric acid. In both studies, results obtained indicated that the degree of chromium removal was proportional to the dosage of the adsorbent used and their contact time. All the works done by these researchers show that biosorption utilizing microorganism and agricultural materials offer an ideal alternative for the treatment of heavy metal polluted water. 7. MICROORGANISMS IN BIOREMEDIATION The ability of microorganisms to utilise natural and synthetic pollutants as substrate for growth is a very important quality upon which bioremediation is based. A lot of work is still ongoing in the area of isolation, identification and characterisation of microorganisms and their potential for bioremediation. Reports suggests that more work still needs to be done to explore microbial diversity with a view to identifying microorganisms with specific and unique qualities vital to bioremediation. Microorganisms indigenous to the site of contamination have been utilised in various bioremediation processes. Information on microbial populations relevant to bioremediation is building up at a fast pace as a result of recent advances in molecular microbial ecology (Watanabe, 2001). This has made available new tools that makes it possible to carry out molecular analyses of microbial populations at contaminated and bioiremediation sites. Microorganisms can be isolated from virtually any environmental condition as they are able to adapt even in very extreme conditions of temperature, oxygen, water, pH etc. The major requirement for growth is an energy and a carbon source. The ability of microorganisms to adapt is what makes them very versatile in the bioremediation of contaminated sites

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Amenaghawon and Obahiagbon Wastewater Treatment by Bioremediation Methods

(Sharma, 2012). These microorganisms can be divided into the following groups: (a) Aerobic microorganisms or aerobes carry out metabolic activities in the presence of oxygen. They require oxygen to oxidise susbtrates through cellular respiration. Examples of aerobic microorganisms with the capacity for biodegradation include Pseudomonas, Alcaligenes, Sphingomonas, Rhodococcus, and Mycobacterium (Giavasis et al., 2006). These microorganism have been reported to possess the capacity to degrade organic pollutants such as aliphatic hydrocarbons, polycyclic aromatic hydrocarbons, pesticides etc. (Vidali, 2001). (b) Anaerobic microorganisms can carry out metabolic activities in the absence of oxygen. They are not as common in use compared to aerobic microorganisms. However, there is an increasing interest in the use of anaerobic microorganisms for the biodegradation of polychlorinated biphenyls (PCBs) in river sediments and the dechlorination of the solvent like trichloroethylene (TCE), and chloroform (Sharma, 2012). (c) Ligninolytic fungi such as white rot fungus like Phanaerochaete chrysosporium have been reported to have the ability to degrade an extremely diverse range of recalcitrant and toxic contaminants. (Adenipekun and Fasidi, 2005). (d) Methylotrophs are aerobic bacteria that utilise methane for metabolic activities. They have the ability to degrade a wide range of organic contaminants such as chlorinated aliphatic trichloroethylene and 1,2dichloroethane. For efficient biodegradation, it is important that the microorganism and the target contaminant be in intimate contact. . This can be enhanced by making use of some surfactants such as sodium dodecyl sulphate (SDS).

carbon, nitrogen and phosphorus and some others in lesser amounts but carbon is needed in greater proportions than the others. These nutrients are often present in wastewater stream but not in the proportion required by the cells for optimum metabolic activities. The lack of nitrogen and phosphorus limits the rate of biodegradation. In the light of this, it becomes important to ensure adequate supply of these important nutrients to enhance biodegradation rates. This is usually accomplished through biostimulation which involves the addition of limiting nutrient such as nitrogen and phosphorus to the wastewater stream. Biostimulation has been reported to enhance the biodegradation of organic pollutants (Obahiagbon et al., 2009; Otokunefor and Obiukwu, 2010)

8. FACTORS OF BIOREMEDIATION

8.3. Temperature

A host of factors can affect the extent and effectiveness of bioremediation. These factors are environmental in consideration and include the availability of nutrients and oxygen. These can be readily manipulated using effective biostimulation strategies. Other factors include temperature and pH of the remediation medium. These however are not easily controllable.

Temperature is another important environmental factor that affects the rate of bioremediation. In the same way that chemical reactions are affected by temperature, biochemical reactions upon which the process of bioremediation is based are also temperature dependent. A temperature increase results in a decrease in viscosity of liquid organic pollutants, consequently affecting the degree of distribution and increasing diffusion rates of the compounds. Typically, an increase in temperature favours the biodegradation reaction. However, above a certain optimum temperature which is organism specific, the activity of the microorganism begins to slow and they subsequently die. Hence it is important to identify this optimum and ensure that bioremediation operations are maintained at that temperature.

8.2. Oxygen This is one of the most important requirements for microbial degradation. Most wastewater treatment facilities adopt aeration based treatment strategies. In such cases, the availability of oxygen becomes a critical factor. Oxygen is generally necessary for the initial degradation of oil, and subsequent reactions may also require direct incorporation of oxygen. Typically, 3 to 4 parts of dissolved oxygen are necessary to completely oxidize 1 part of oil into carbon dioxide and water (Giavasis et al., 2006). Though anaerobic degradation of oil in wastewater can occur, it is however in very small degrees. For oil spills on the ocean surface, oxygen is not usually a factor that limits the rate of biodegradation as there is plentiful supply of oxygen close to the surface of the ocean. However, inadequate supply of oxygen limits the extent of biodegradation. This is the reason why it takes longer to degrade oil that has sunk below the surface of the water.

8.1. Nutrients Microorganisms need nutrients to survive. These nutrients are the basic building blocks of living things and enable microorganisms to carry out metabolic activities needed for the breakdown of contaminants during bioremediation. All microorganisms need

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Babarinde A, Babalola JO, Adegoke J, Osundeko AO, Ibidapo TJ, Nwabugwu CA, Ogundimu OF (2012). Biosorption of Ni(II), Cr(III), and Co(II) from Aqueous Solutions using Cocoyam (Colocasia esculenta) Leaf: Kinetic, Equilibrium, and Thermodynamic Studies. Pacific Journal of Science and Technology. 13(2): 272-282. Bamforth SM, Singleton I (2005). Bioremediation of polycyclic aromatic hydrocarbons: current knowledge and future directions. Journal of chemical technology and biotechnology, 80: 723-736. Bishnoi NR, Bajaj M, Sharma N, Gupta A (2004). Adsorption of Cr(vi) on activated rice husk carbon and activated alumina. Bioresource Technology, 91: 305–307. Bradley PM, Landmeyer JE (2006). Low-temperature MTBE biodegradation in aquifer sediments with a history of low, seasonal ground water temperatures. Ground Water Monitoring and Remediation, 26(1): 101–105. Cauwenberghe LV, Roote DS (1998). In situ bioremediation. Technology overview report, Ground water remediation technologies analysis center, p.4. Chaillana F, Flècheb A, Burya E, Phantavonga Y-hui, Saliot A, Oudot J (2004). Identification and biodegradation potential of tropical aerobic hydrocarbon-degrading microorganisms. Research in Microbiology. 155(7): 587-595. Chen C, Wang J (2008). Removal of Pb2+, Ag+, Cs+ and Sr2+ from aqueous solution by brewery’s waste biomass. Journal of hazardous materials, 151: 65-70. Congeevaram S, Dhanarani S, Park J, Dexilin M, Thamaraiselvi K (2007). Biosorption of chromium and nickel by heavy metal resistant fungal and bacterial isolates. Journal of hazardous materials, 146: 270-277. Cosgrove L, McGeechan PL, Handley PS, Robson GD (2010). Effect of biostimulation and bioaugmentation on degradation of polyurethane buried in soil. Applied and environmental microbiology, 76(3): 810-819. Costley SC, Wallis FM (2001). Bioremediation of heavy metals in a synthetic wastewater using a rotating biological contactor. Water research, 35(15): 3715-3723. Crawford RL, Crawford DL Eds. (2005). Bioremediation: Principles and Applications, Biotechnology Research Series: 6, Cambridge University Press, UK. Cunningham JA, Rahme H, Hopkins GD, Lebron C, Reinhard M (2001). Enhanced in situ bioremediation of BTEX-contaminated

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International Biodeterioration and Biodegradation Journal, 59: 40-43. Hedlund BP, Staley JT (2001). Vibrio cyclotrophicus sp. nov., a polycyclic aromatic hydrocarbon (PAH)-degrading marine bacterium. International journal of systematic and evolutionary microbiology, 51(1): 61-66. Hidayat A, Tachibana S (2012). Bioremediation of Aliphatic Hydrocarbon in Three Types of Crude Oil by Fusarium sp. F092 under Stress with Artificial Sea Water. Journal of Environmental Science and Technology, 5(1): 64-73. Ho YS (2005). Effect of pH on lead removal from water using tree fern as the sorbent. Bioresource Technology, 96: 1292–1296. Janbandhu A, Fulekar MH (2011). Biodegradation of phenanthrene using adapted microbial consortium isolated from petrochemical contaminated environment. Journal of hazardous materials, 187: 333-340. Jia C, Batterman S (2010). A Critical Review of Naphthalene Sources and Exposures Relevant to Indoor and Outdoor Air. International Journal of Environmental Research and Public Health, 7: 2903-2939. Kadirvelu K, Senthilkumar P, Thamaraiselvi K, Subburam V (2002). Activated carbon prepared from biomass as adsorbent: elimination of Ni(II) from aqueous solution. Bioresource Technology 81: 87–90. Kao CM, Huang WY, Chang LJ, Chen TY, Chien HY, Hou F (2006). Application of monitored natural attenuation to remediate a petroleumhydrocarbon spill site. Water Science and Technology, 53: 321–328. Kim SJ, Choi DH, Sim DS, Oh YS (2005). Evaluation of bioremediation effectiveness on crude oilcontaminated sand. Chemosphere, 59(6): 845852. Khan K, Naeem M, Arshed MJ, Asif M (2006). Extraction and characterization of oil degrading bacteria. Journal of Applied Sciences, 6: 23022306. Kshirsagar AD (2013). Bioremediation of wastewater by using microalgae: an experimental study. International Journal of Life Sciences Biotechnology and Pharma Research, 2(3): 339346. Lendvay JM, Löffler FE, Dollhopf M, Aiello MR, Daniels G, Fathepure BZ, Adriaens P (2003). Bioreactive barriers: a comparison of bioaugmentation and biostimulation for chlorinated solvent remediation. Environmental Science & Technology, 37(7): 1422-1431. Li PF, Mao ZY, Rao SJ, Wang XM, Min MZ, Qiu LW, Liu ZL (2004). Biosorption of uranium by

lake harvested biomass from cyanobacterium bloom. Bioresource Technology, 94: 193–195 Li Q, Kang C, Zhang C (2005). Waste water produced from an oilfield and continuous treatment with an oil-degrading bacterium. Process Biochemistry, 40(2): 873-877. Li Z, Mulholland JA, Romanoff LC, Pittman EN, Trinidad DA, Lewin MD, Sjödin A (2010). Assessment of non-occupational exposure to polycyclic aromatic hydrocarbons through personal air sampling and urinary biomonitoring. Journal of Environmental Monitoring, 12(5): 1110-1118. Lin C, Gan L, Chen ZL (2010). Biodegradation of naphthalene by strain Bacillus fusiformis (BFN). Journal of hazardous materials, 182: 771-777. Lodeiro P, Cordero B, Barriadh JL, Herrero R, deVicente MES (2005). Biosorption of cadmium by biomass of brown marine macroalgae. Bioresource Technology, 96: 1796– 1803. Macek T, Mackova M, Kas J (2000). Exploitation of plants for the removal of organics in environmental remediation. Biotechnology Advances, 18: 23-34. Margesin R, Schinner F (2001). Bioremediation (Natural Attenuation and Biostimulation) of Diesel-Oil-Contaminated Soil in an Alpine Glacier Skiing Area. Applied and Environmental Microbiology, 67(7): 3127-3133. Morill JA, Antizar-Ladislao B, Monteoliva-Sanchez M, Ramos-Cormenzana A, Russell NJ (2009). Bioremediation and biovalorisation of olive-mill wastes. Applied Microbiology and Biotechnology, 82: 25-39. Mrozik A, Piotrowska-Seget Z (2010). Bioaugmentation as a strategy for cleaning up of soils contaminated with aromatic compounds. Microbiological Research, 165(5): 363-375. Mukred AM, Hamid AA, Hamzah A, Yusoff WMW (2008). Development of three bacteria consortium for the bioremediation of crude petroleum oil in contaminated water. Online Journal of Biological Sciences. 8(4): 73-79. Newman LA, Reynolds CM (2004). Phytodegradation of organic compounds. Current Opinion in Biotechnology, 15: 225-230. Nikolopoulou M, Kalogerakis N (2009). Biostimulation strategies for fresh and chronically polluted marine environments with petroleum hydrocarbons. Journal of Chemical Technology and Biotechnology, 84(6): 802-807. NTP (2004). Report on Carcinogens - Naphthalene, Eleventh Edition; U.S. Department of Health

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amended with biochar. International Biodeterioration & Biodegradation, 85: 150155. Röling WF, Milner MG, Jones DM, Lee K, Daniel F, Swannell RJ, Head IM (2002). Robust hydrocarbon degradation and dynamics of bacterial communities during nutrient-enhanced oil spill bioremediation. Applied and Environmental Microbiology, 68(11): 55375548. Samanta SK, Singh OV, Jain RK (2002). Polycyclic aromatic hydrocarbons: environmental pollution and bioremediation. Trends in biotechnology, 20(6): 243-248. San Miguel V, Peinado C, Catalina F, Abrusci C (2009). Bioremediation of naphthalene in water by Sphingomonas paucimobilis using new biodegradable surfactants based on poly (ɛcaprolactone). International Biodeterioration & Biodegradation, 63(2): 217-223. Schiewer S, Patil SB (2008). Modeling the Effect of pH on Biosorption of Heavy Metals by Citrus Peels. Journal of Hazardous Materials. 157:817. Schmidt LM, Delfino JJ, Preston JF, Laurent GS (1999). Biodegradation of low aqueous concentration pentachlorophenol (PCP) contaminated groundwater. Chemosphere, 38: 2897–2912. Sharma S (2012). Bioremediation: Features, Strategies and applications. Asian Journal of Pharmacy and Life Science, 2(2): 202-213. Smith AE, Hristova K, Wood I, Mackay DM, Lory E, Lorenzana D, Scow KM (2005). Comparison of biostimulation versus bioaugmentation with bacterial strain PM1 for treatment of groundwater contaminated with methyl tertiary butyl ether (MTBE). Environmental health perspectives, 113(3): 317-322. Srivastava NK, Majumder CB (2008). Novel biofiltration methods for the treatment of heavy metals from industrial wastewater. Journal of hazardous materials, 151: 1-8. Subramanian M, Oliver DJ, Shanks JV (2006). TNT Phytotransformation Pathway Characteristics in Arabidopsis: Role of Aromatic Hydroxylamines. Biotechnology Programme, 22: 208 -216. Thavasi R, Jayalakshmi S, Balasubramanian T, Banat IM (2006). Biodegradation of crude oil by nitrogen fixing marine bacteria Azotobacter chroococcum. Research Journal of Microbiology 1(5): 401-408. Ting WTE, Yuan SY, Wu SD, Chang BV (2011). Biodegradation of phenanthrene and pyrene by Ganoderma lucidum. International

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Biodeterioration & Biodegradation, 65(1): 238242. Thomassin-Lacroix E, Eriksson M, Reimer K, Mohn W (2002). Biostimulation and bioaugmentation for on-site treatment of weathered diesel fuel in Arctic soil. Applied microbiology and biotechnology, 59(4-5): 551-556. Thompson IP, Van Der Gast CJ, Ciric L, Singer AC (2005). Bioaugmentation for bioremediation: the challenge of strain selection. Environmental Microbiology, 7(7): 909-915. Verma P, George K, Singh H, Singh S, Juwarkar A, Singh R (2006). Modeling rhizofiltration: heavy metal uptake by plant roots. Environmental Modeling and Assessment, 11: 387-394. Vidali M (2001). Bioremediation. An overview. Pure and Applied Chemistry, 73(7): 1163-1172. Vogel TM (1996). Bioaugmentation as a soil bioremediation approach. Current Opinion and Biotechnology 7: 311–316. Wang JL, Chen C (2006). Biosorption of heavy metal by Saccharomyces cerevisiae: a review. Biotechnology Advances, 24: 427–451. Watanabe K (2001). Microorganisms relevant to bioremediation. Current opinions in biotechnology, 12: 237-241. Whitehead PG, Hall G, Neal C, Prior H (2005). Chemical behaviour of the Wheal Jane bioremediation system. Science of the Total Environment, 338: 41–51. Widdowson MA, (2004). Modeling natural attenuation of chlorinated ethenes under

spatially varying redox conditions. Biodegradation, 15: 435–451. Woo OT, Chung WK, Wong KH, Chow AT, Wong PK (2009). Photocatalytic oxidation of polycyclic aromatic hydrocarbons: Intermediates identification and toxicity testing. Journal of hazardous materials, 168(2): 1192-1199. Wilson NK, Chuang JC, Lyu C, Menton R, Morgan MK (2003). Aggregate exposures of nine preschool children to persistent organic pollutants at day care and at home. Journal of Exposure Science and Environmental Epidemiology, 13(3): 187-202. Yerushalmi L, Manuel MF, Guiot SR (1999). Biodegradation of gasoline and BTEX in a microaerophilic biobarrier. Biodegradation, 10: 341–352. Yu KSH, Wong AHY, Yau KWY, Wong YS, Tam NFY (2005). Natural attenuation, biostimulation and bioaugmentation on biodegradation of polycyclic aromatic hydrocarbons (PAHs) in mangrove sediments. Marine pollution bulletin, 51(8): 1071-1077. Zhuang P, Yang QW, Wang HB, Shu WS (2007). Phytoextraction of heavy metals by eight plant species in the field.Water. Air and Soil Pollution, 184: 235-242. Zouboulish AI, Loukidou M, Matis KA (2004). Biosorption of Toxic Metals from Aqueous Solutions by Bacteria Strains Isolated from Metal Polluted Soils. Process Biochemistry, 39: 909-916.

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Supplementation of Novel Solid Carbon Source Prepared from Dried AttachedGrowth Biomass for Bioremediation of Wastewater Containing Nitrogen Jun-Wei Lim, Mohammed J.K. Bashir*, Choon-Aun Ng, Xinxin Guo Department of Environmental Engineering, Faculty of Engineering and Green Technology (FEGT), University Tunku Abdul Rahman, 31900 Kampar, Perak, Malaysia *Corresponding Author: [email protected]; Tel: 605-4688888 ext: 4559; Fax: 605-4667449 Abstract. The main objective of this study is to validate the feasibility of using dried attached-growth biomass from the polyurethane (PU) foam cubes as a solid carbon source for the enhancement of denitrification process in the intermittently aerated moving bed sequencing batch reactor (IA-MBSBR). The IA-MBSBR packed with PU foam cubes coated with dried attached-growth biomass could maintain approximately 80% of total nitrogen (TN) removal efficiency for 8 consecutive cycles of operation. Subsequently, the exhausted stored carbon source within the PU foam cubes could be replenished by merely drying the fresh attached-growth biomass formed when the cubes were used as a carbon source. Thus, the reuse/recycle of biomass-coated PU foam cubes is possible, making it a sustainable solid carbon source for the enhancement of denitrification process in bioremediating wastewater containing nitrogen-cum-low COD/N via IA-MBSBR. Keywords: Bioremediation, Wastewater, Solid Carbon Source

improper sewage treatment and disposal. In some industrial activities such as fossil fuels combustion and fertilizers, explosive, glass, plastics and foods productions also can contribute to nitrate pollution (Robinson-Lora and Brennan, 2009; Wang et al., 2013). In nature, the concentrations of nitrate in groundwater are usually less than 2 mg/L (Mueller et al., 1995). However, in contaminated areas, nitrate concentrations can exceed 200 mg/L (ITRC, 2002). The Water Quality Assessment program of the US Geological Survey reports that nitrate is the pollutant that most frequently exceeds its standard limits (Squillace et al., 2002). The main health effect associates to the ingestion of water contaminated with high concentration of nitrate are the occurrence of methemoglobinemia notably in infants or “blue-baby syndrome”. Some studies have demonstrated that nitrate can be endogenously reduced to nitrite, which can then undergo nitrosation reactions in the stomach with amines and amides to form various N-nitroso compounds, most of which are extremely carcinogenic (WHO, 2004; Yang et al., 2007). On that account, appropriate standards have been set by various agencies. The USEPA (2000) has set the maximum contaminant levels of 10 and 1 mg/L for nitrate-nitrogen (NO3--N) and nitrite-nitrogen (NO2-N), respectively, in drinking water. The World Health Organization and European Economic Community have set the standards of 11.3 mg/L for NO3--N which are later adopted as national standard for drinking water by most of the countries in the world (Wang, 2013). Therefore, for the sake of fulfilling the standards requirement, the discharge wastewater containing nitrate must be stringently treated before releasing to the environment in order to minimize the possibilities of contamination of potable water with nitrate.

1. INTRODUCTION It is undeniable that water is an indispensable element for survival of all living creatures. Although the Earth's surface is virtually covered with 70% of water, its presence is not limitless. Of all the water, 97% consist of salt water which is unacceptable for the direct human consumption. To top it off, of the remaining 3% of potable water, only about 15% is easily accessible, e.g., rivers, streams, creeks, ponds, etc., and about 85% is found in ice floes and glaciers, neither of which are readily accessible (Sills, 2003). Owing to the tremendous increase of human growth in recent decades, water scarcity has emerged as one of the dire issue for communities across the country. In United States, almost all the region of the country has experienced water shortages in the last five years (USEPA, 2008). As per United Nations, every day approximately 4400 children under the age of five die because of diseases cause by contaminated water ingestion and sanitation (Ghaitidak and Yadav, 2013). In a third world’s population, one in every six persons has no access to clean water within a kilometer of reach (Ghaitidak and Yadav, 2013). In parts of Asia and Africa, it was estimated that the people under a threshold of water stress, i.e., accessible of renewable water resources 11, a gradient of chloride ions from the strip to the feed phase provides a driving force for the mass transport. According to Wieczorek et al. (1997), on their investigation on concentration of amino acids using supported liquid membranes with di-2ethylhexyl phosphoric acid as a carrier, the extractions are made from an aqueous donor phase with pH 3 to a more acidic acceptor phase and the mass transfer is driven by the proton gradient between these phases.

D = kT/(6μ η r) (11) where D is diffusion coefficient (cm2/s), k is Boltzmann constant, T is absolute temperature (Kelvin), η is the organic phase viscosity and r is the pore radius of the solute. 4.1. Permeation in SLM Permeation is the process in which a solute dissolves in the organic phase and diffuses towards the stripping phase. Substance solubility plays the major role in simple transport permeation and equilibrium level is the final step of permeation. In other words, the transported compound is in the same condition either in feed phase or stripping phase (Cussler et al., 1989). The advantage of SLM system is that by using facilitated organic phase, the transport of solute through the LM and as a consequence solute permeability and selectivity can be enhanced drastically. In this method, the reaction between solute and the selected carrier only takes place at the membrane or in better words at the interface of feed phase and membrane (Juang et al., 1998). If C is the carrier in the LM which is able to form a complex with the solute S: S + C ↔ S=C (12) then the dissociation constant is:

4.3. Organic Solvents Selection in SLM Hydrophobicity is the main characteristic for an organic solvent to be used in SLM to ensure immiscibility with aqueous feed and stripping phases. Low viscosity of an organic solvent also plays a major role in mass transfer of solute through the LM but at the same time it has a negative effect on SLM stability. Low volatility, low interfacial tension between aqueous and organic phases in the support pores which lead to the higher mass transfer can be considered as important parameters for organic solvent selection in SLM. Table 1 Shows selected parameters of the commonly used organic solvents in SLM (Rydberg et al., 2004).

Ka = [S=C]/[S][C] (13) From the Fick’s first law, the flux is:

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Table 1: selected parameters of the commonly used organic solvents in SLM (Rydberg et al., 2004) Organic Solvent Dodecane Heptane Toluene Kerosene Diphenyl methane Diheyl ether 1-octanol

Density x 10-3 (kg/m3) 0.75 0.68 0.78 0.79 1.00 0.79 0.83

Viscosity x 103 (Pa s) 1.50 0.38 0.54 1.24 2.96 1.87 7.47

Surface tension x 10 3 (N/m) 24.9 19.6 27.9 25.3 38.4 27.1

Solubility in water x 10-3 (kg/m3) 0.07 6.51 -

A good carrier that can be used to enhance the selectivity should have the following characteristics: (i). Formation and decomposition of the complex on membrane interfaces should be fast and rapid. (ii). Side reactions can decrease the selectivity and extraction process. (iii). Irreversible reaction and degradation are considered as limiting parameters in selectivity. (iv). Low solubility in the aqueous feed and strip phases of the carrier has a key role for choosing a suitable carrier. (v). Should not be hazardous or toxic to the environment and should be cost effective specially in industrial applications. As the ionic carriers, mostly amines or carboxylic and phosphoric acids for metals, organic acids and amines are typically used. For metal extraction, the addition of thiocyanate ions to the donor is needed to form a negatively charged metal thiocyanate complex, which can give an ion pair with the carrier (Papantoni et al., 1995). The other parameter that can affect the selectivity is the diffusion coefficient which depends on the molecular radius of the solute. Changing pH in feed and stripping phases for the purpose of activating and deactivating the compounds in these phases can increase the selectivity as well. For example, the basic feed phase and acidic strip phase is useful for selective amines extraction (Dzygiel and Wieczorek, 2010). If the pH of the feed (donor) phase is adjusted to a sufficiently high value, the transported amines are uncharged and are transported over the organic liquid that is used as a membrane phase. The strip (acceptor) phase on the other side of the membrane is an acidic solution or buffer with low pH.

4.4. SLM and Species Recovery One of the most important issues in extraction and separation process is substance recovery which is directly related to pertraction efficiency and/or recovery. The solute concentration difference ΔS over the membrane can be expressed as: ΔS = αfSf - αsSsKs/Kfm (17) where αf and αs as are the fractions of the transported substance, which are pertractable from the feed to the strip phases, respectively. Ksm is the partition coefficient for the solute between the strip and membrane phase, and Kfm is the partition coefficient for the solute between the feed and membrane phase. While the feed and strip phases are mostly aqueous, both partition coefficients are similar (Kislic 2010). Other parameters like diffusion coefficients in feed and membrane phases, partition coefficients or membrane thickness can also affect the extraction rate. According to Jonsson and Mathiasson (1999), there are two possibilities for pertraction controlling: (i). Membrane controlled pertraction, when there is a limiting step for diffusion of the transported solute through the liquid membrane. (ii). Feed controlled pertraction, when the diffusion through the feed phase to the feed-membrane interface appears as the limiting step. 4.5. SLM and Selective Extraction The selective extraction in SLM is the ability to transfer the desired compounds only and not the interfering or unwanted compounds. The selectivity depends mostly on the species capture method and also the used transport mechanism (Thornton, 1992). When the simple permeation is applied, the selectivity is not high and is governed by solubility differences between the solutes in the membrane phase; however, when carrier is used the transport efficiency and consequently the selectivity increases. Various carrier molecules or ions can be incorporated in the membrane phase to enhance the selectivity and mass transfer (Dzygiel and Wieczorek, 2010).

4.6. SLM Unit Design Pertraction, extraction and transport processes in SLM like all LM types, highly depend on the membrane design and constructor. Microporous polymeric membrane is typically used for the membrane phase in SLM design and modification. The type of polymeric microporous membrane has a direct impact on the membrane lifetime, stability, performance and efficiency. Nowadays, the new generation of developed inorganic

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membranes like ceramics, porous metals and zeolites are used in SLM reactor modification and design and have shown a range of advantages like thermal and mechanical stability, being resistant to chemical and organic solvents and being recyclable. The utilization of advanced inorganic membrane materials is nowadays very important. Cot et al. (2000) worked on preparation of inorganic membrane materials innovative concepts like templating effect, nanophase materials, growing of continuous zeolite layers, and hybrid organic–inorganic materials with permselective properties for gas separation and facilitated transport of solutes in liquid media which have been successfully adapted to membrane applications. Resina et al. (2008) used hybrid and activated composite membranes containing Aliquat 336 for the transport of Pt(IV) and McCleskey et al. (2002) used a thin layer of gold (700 Å) and have reported high selectivity of U over Eu until [U] is 6

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According to the result obtained, removal of bacteria which was higher than 6 logs (totally removed) was achieved by the PES 13 wt.% membrane even high concentration was used as feed solution. Changing pressure did not post significant effect on change of the removal ability of the membrane. However, increasing pressure during application enhance significantly the flux production. As totally rejection obtained, it means that produced UF membrane had a smaller pores size than the bacteria. According to Tsummi et al. (1990), in the cell separation by such porous membrane, molecular size exclusion effect is dominant factor. When the size of the solute is much larger than pore size of the membrane, effective cutoff occurs. Once it achieved totally rejection to the bacteria, it shows that produced membrane was safe to be used for bacteria filtration process as it was impossible for bacteria which have bigger size than the membrane pores can pass through it.

problems. Generally a membrane with bigger pore size and loose structure will have higher flux, but lower rejection ability. By increasing polymer concentration in the solution, it will make the membrane become denser and usually rejection ability will be improved as well, but flux production will be decreased. This is supported by Li (1993) where in his study, he found out that as flux increase, membrane selectivity will decrease and as the selectivity increased, fluxes will be decreased. Even though there is no perfect membrane, which can serve all application with perfect rejection and high flux production, but still there is optimum membrane which can serve in specific application with the acceptable range of removal ability and flux production. 7. CONCLUSION Based on the experimental data, this study had proven that membrane with ternary system and concentration of PES 13.60 wt.% was the best among the produced membranes. It showed excellent performance in terms of flux (2700 L/m2.h) and achieved 99.97% for salt rejection (NaCl 0.01 M). In addition, the membrane also achieved 100% rejection for the bacteria (E. coli and E. faecalis) with an average flux of 320 L/m2.h and 400 L/m2.h at low pressure provided (1 bar). SEM cross sectional images clearly observed that this asymmetric membrane provides great properties and excellent separation performances. As the produced membrane able to totally remove the bacteria in water, this may help to provide a safe and clean drinking water for those needed.

6. CHALLENGES IN FABRICATING SUITABLE MEMBRANE FOR WATER TREATMENT APPLICATION There are several parameters must be considered in order to produce high performance membranes, such as dope formulation, casting condition, ambient temperature during membrane casting, shear rate, membrane thickness and so on. Changes on each parameter, may affect membrane properties, and hence changing the membrane performance in different application. In order to reproduce a high performance membrane, all the parameters involved during membrane fabrication must be recorded in detail. Another challenge is to produce membranes which give good fluxes and rejection. These always become a limitation when it comes to the flux and selectivity

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Lee N, Amy G, Philippe J, Brisson CH (2004). Identification and understanding of fouling in low-pressure membrane (MF/UF) filtration by natural organic matter (NOM). Water Research, 38: 4511-4523. Liu C (2014). Advances in Membrane Technologies for Drinking Water Purification. Comprehensive Water Quality and Purification. Vol 2, 75-97. Lo YM, Yang ST, Min DB (1996). Kinetic and feasibility studies of ultrafiltration of viscous xanthan gum fermentation broth. Journal of Membrane Science, 117 (1–2): 237–249. Mark CP (1988). Handbook of industrial membrane technology. Noyes Publication. Park Ridge, New Jersey, U-S-A. Monarca S, Zani C, Richardson S D, Thruston Jr A D, Moretti M, Feretti D, Villarini M (2004). A New Approach to Evaluating the Toxicity and Genotoxicity of Disinfected Drinking Water. Water Research, 38: 3809-3819. Morita RY (1993). Bioavailability of energy and the starvation state. In: Starvation in Bacteria (ed. Kjelleberg. ) S, pp. 1 23. New York: Plenum Press. Mulder M (1996). Basic principles of membrane technology. Dordrecht, Kluwer Academic Publishers, Netherlands. Nunes SP, Peinemann KV (2001). Introduction, in Membrane Technology: in the Chemical Industry (eds S. P. Nunes and K.-V. Peinemann), Wiley-VCH Verlag GmbH, Weinheim, FRG. Rook J J (1974). Formation of haloforms during chlorination of natural waters. Journal of Water Treatment Examine, 23: 234–243. Sadr Ghayeni SB, Beatson PJ, Fane AJ, Schneider RP (1999). Bacteria Passage Through Microfiltration Membranes In Wastewater Applications. Journal of Membrane Science, 153: 71-82. Tsummi T, Sato T, Osawa N, Hitaka H, Hirosawa T, Yamaguch K, Hamamoto Y, Manabe S, Yamasiki T, Yamamoto N (1990). Structure and filtration performance of improved cuprammonium regenerated cellulose hollow fiber (Improved BMM hollow fiber) for virus removal. Polymer Journal, 22: 1085-1100. Xia S, Nan J, Liu R, Li G (2004). Study of Drinking Water Treatment by Ultrafiltration of Surface Water and its Application to China. Desalination, 170: 41-47.

ACKNOWLEDGEMENTS We would like to extend our gratitude to Ministry of Science, Technology and Innovation (MOSTI) for the fund with project No. 03-02-11-SF0161. REFERENCES Brock TD (1983). Membrane Filtration: A User’s Guide and reference Manual. Springer, New York. Cheryan M (1998). Ultrafiltration and microfiltration handbook. Technomic Publishing, Lancaster, 2nd Edition, Florida. Choi K Y J, Dempsey BA (2004). In-line coagulation with low-pressure membrane filtration. Water Research, 38: 4271–4281. Drinking Water Contaminants (2006). Basic information about E. coli O157:H7 in drinking water. http://www.epa.gov/safewater/contaminants/eco li.html. Fane AG, Yeo A, Law A, Parameshwaran K, Wicaksana F, Chen V (2005). Low pressure membrane processes ~ doing more with less energy. Desalination, 185: 159-165. Gupta V K, Ali I (2013). Water Treatment by Membrane Filtration Techniques. Environmental Water. 135-154. Li S G (1993). Preparation of Hollow Fiber Membranes for Gas Separation. Enschede. Netherlands. 1-22 Hasan A A, Victor K, Nidal H (2013). Hybrid ion exchange – Pressure driven membrane processes in water treatment: A review. Vol 116, 253–264. Hendricks D (2005). Water Treatment Unit Processes Physical and Chemical. Taylor & Francis Group, Boca Raton, 913-915. Ismail AF, Hassan AR (2005). Formation and Characterization of Asymmetric Nanofiltration Membrane: Effect of Shear Rate and Polymer Concentration. Journal of Membrane Science, 270: 57 – 72. Ismail AF, Lai PY (2004). Development of defect-free asymmetric polysulfone membranes for gas separation using response surface methodology. Separation and Purification Technology, Vol. 40 (2): 191-207. Ismail AF, Rahman AR (2004). The deduction of fine structural details of asymmetric nanofiltration membrane using theoretical models. Journal of Membrane Science, 231: 25-36.

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Removal of Copper from Aqueous Solution by Using Thermo-responsive Polymer Hydrogel as Copper Chelator via Polymer-enhanced Ultrafiltration (PEUF) J.J. Chen, A.L. Ahmad, B.S. Ooi* School of Chemical Engineering, Engineering Campus, Universiti Sains Malaysia, Seri Ampangan, Nibong Tebal 14300 Pulau Pinang, Malaysia, Tel: +6045996418; Fax :+6045941013; * Corresponding Author: [email protected] Abstract. Membrane based separation is another efficient method for separating inorganic substances from aqueous effluents due to its simplicity to scale-up and low energy consumption. Nevertheless, application of this method to remove heavy metals from wastewaters is still a challenge to most researchers owing to the fact that heavy metals exist in aqueous solution as hydrated ions with low molecular weights that can easily pass through most membranes except membranes for reverse osmosis (RO) and nanofiltration (NF). However, RO and NF membranes with denser structures and narrower pores have very low selectivity where nearly all types of solutes in the effluent are retained instead of separating only the targeted metal ions. Furthermore, high energy is consumed and high operating pressure is required to achieve adequate permeation of solvent through these types of membrane. Advanced materials combined with smart membrane systems could provide low energy and highly sustainable metal recovery process. For example, polymer enhanced ultrafiltration combined with thermo-responsive hydrogel could offer such attractive process. Compared to the conventional adsorbents, the adsorption capacity of the thermoresponsive hydrogel is a function of solution pH, ligand density as well as temperature. The separation of hydrogel from the waste stream can be realized through temperature induced agglomeration as well as steric retention by the membrane. The separation process could be almost instantaneous with the metal recovery obtained even at very trace concentrations. Lower temperature is more favorable due to the swollen hydrogel that exerts less diffusional resistance for metal to bind within the interior marix. Low temperature is also favorable as it promises less membrane fouling phenomenon. Keywords: Copper removal, Polymer-enhanced Ultrafiltration, Aqueous Solution

on metal removal is currently resolved by integration of both metal-polymer complexation and UF (Micheals, 1990). This combination is known as polymer-enhanced ultrafiltration (PEUF). In PEUF operation, water-soluble polymer with metal-chelating ability is used as chelating agent to complex with a targeted metal ion. The metal ion is chelated by functional ligands of the polymer such as carboxylic (-COOH) and amide (-CONH) groups via electrostatic interaction to form metal-polymer complexes. The complex can then be retained by UF membrane effectively (Almutairi and Lovitt, 2012, Camarillo et al., 2012, Zerze et al., 2013). Complexation of metal in aqueous solution is deemed as a major type of metal ion adsorption by polymeric substances through chelation process. Polymer with metal-chelating ligands are known as chelating agent (chelator) that bind with free metal ions in aqueous solution to form complex molecules known as ‘chelates’. A chelator contains one or more ligands that are ionic or neutral such as –SO3H, -COOH, -NH2, -OPO3H and -C=O. Each ligand could possess mono-, bi- or even polydentate enabling it to form ring structure with a metal ion (Flora and Pachauri, 2010, Butvin et al., 1988).

1. INTRODUCTION Heavy metals are inorganic pollutants commonly found in wastewaters discharged from industries like metal-plating, tanneries, batteries, automobile and fertilizer manufacturing. The pollution of natural water-bodies especially rivers by the metalcontaminated industrial effluents is becoming one of the most serious environmental problems. Several conventional methods had been widely adopted to scavenge metal ions from aqueous solution including ion exchange resins, electrodialysis, chemical precipitation, adsorption and membrane separation. However, the efficiency of these methods has been constraint by some unsolved challenges such as production of secondary sludge, low selectivity and low metal-recovery efficiency (Lazaridis et al., 2004, Dermentzis et al., 2009, Ghodbane et al., 2008). In terms of low energy consumption and high solute-rejection efficiency, ultrafiltration (UF) is deemed as a promising membrane separation process for the treatment of industrial wastewaters. The average pore diameters of UF membranes fall within the range of 10 – 1000 Ǻ (1 – 100 nm). The pore structures at this range of pore diameters allow permeation of microsolutes with molecular weight (MW) less than 300 g/mol. Nevertheless, copper divalent ion (Cu2+) as a transition metal ion with molar mass of merely 63.55 g is too low to be retained by a UF membrane. The shortcoming of UF process

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protonated ligands, [LH], deprotonated ligands [L] and chelated ligands [MLn] is expressed as: (1.4) where [L]total is the total ligand concentration in the solution. In addition, the extent of metal-ligand complexation is described by a complex formation constant, Kf which is defined as: (1.5)

1. COMPLEXATION OF COPPER IONS BY POLYMER WITH METAL-CHELATING LIGANDS The metal-ligand complex formation between a metal ion and ligands of a polymer can be described by the following reaction: (1.0) where M refers to a free metal ion; L is a ligand of polymer, and n is known as the number of ligands involved in chelation with the metal ion. The chelation of metal ion by acidic ligands of a polymer involves dissociation of acidic functional groups like carboxylic groups (-COOH) into both anionic carboxylates (-COO-) and H+ protons at pH higher than pKa of the acidic ligand involved, for example, pKa of acrylic acid (AA) is around 4.3. Hence, it is important to determine their dissociation equilibrium prior to analyzing the copper binding equilibrium. Camarillo et al. (2012) elucidated that the dissociation equilibrium constant, Ka of protonated ligands is expressed as:

where: [MLn] = concentration of metal-ligand complex formed; [M] = concentration of metal ion; [L] = concentration of protonated ligand The value of this formation constant is highly dependent on solution pH, and it is also expected to be influenced by temperature for thermo-responsive PNIPAM-co-AA polymer hydrogels. For ease of experimental measurements and calculation, the chelating ability of a polymer is usually quantified by its chelating capacity in equilibrium state, which is always expressed as a parameter known as equilibrium adsorption capacity, qe in most of the metal adsorption studies as shown in Equation 1.6: (1.6)

(1.1)

where V (L) is the total volume of reaction solution, M (g) is the dry mass of chelating polymer.

where L- is the deprotonated ligands of polymers, whereas HL refers to protonated ligands and β represents a constant depending on each polymer ligand which accounts for nearest neighbouring interaction (Camarillo et al., 2012, Porasso et al., 2000). This expression is also applicable to other polymers with acidic ligands in which their chelating ability is mainly contributed by their acidic functional groups. Since the dissociation of carboxylic groups is dependent on pH, therefore solution pH in the absence of metallic ions is expressed as a function of Ka, β and the degree of protonation, α. The expression is known as modified Henderson-Hasselbach equation as shown in Equation 1.2: (1.2)

2. APPLICATION OF THERMO-RESPONSIVE PNIPAM-co-AA POLYMER HYDROGEL AS NOVEL CHELATING AGENT Selection of a suitable chelating agent is important for optimizing the efficiency of metal ions-complexing in PEUF. Water-soluble polymers are usually chosen as potential metal chelators due to the presence of hydrophilic functional groups that act as ligands for metal chelation such as amide (-CONH), carboxylate (-COO-) and carbonyl (-C=O). These functional ligands are Lewis acids or bases that donate electron pair to fill the unoccupied d-orbital of a metal ion to form a chelated complex (Flora and Pachauri, 2010, Hancock and Martell, 1989). There are numbers of studies reported on the use of different water-soluble polymers as metal chelating agents in PEUF processes. Among them, polyethyleneimine (PEI), poly(vinyl sulfonic acid) (PVSA) and poly(acrylic acid) (PAA) are the most common chelators for the evaluation studies of PEUF. Caῆ izares et al. (2002) applied water-soluble PEI (MW = 25,000 g mol-1) and PAA (MW = 250,000 g mol-1) for recovering Cu2+, Ni2+, Pb2+ and Cd2+ cations. The affinity of these two types of polymer toward metal ions are attributed to their metal-chelating ligands like amines (-NH) groups in PEI and carboxylic (-COOH) in PAA (Caῆ izares et al. 2002). PEI also attracted interest from İslamoğlu et al. (2006)

The protonation degree, α is usually expressed as molar ratio of added base (NaOH) to carboxylic acid (-COOH) of AA moieties in a PNIPAM-co-AA polymer hydrogel: (1.3) where [NaOH] represents molar concentration of NaOH added into the reaction solution; CH is the total molar concentration of carboxylic (-COOH) and carboxylate (-COO-) of hydrogel (Camarillo et al., 2012, Flora and Pachauri, 2010). Hence in the presence of metal ions, the concentration balance of

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to study the effect of ionic strength on its complexation with Cd2+ and Ni2+ in PEUF process. They found that in the absence of NaNO3 salt, the retention value of Cd2+ was higher than that for Ni2+. In contrast, as 0.5 N NaNO3 was added to the binary metal solution, the retention of Ni2+ became higher instead (İslamoğlu et al. 2006). In 2008, Llanos and partners synthesized a novel derivative of PEI known as partially ethoxylated polyethylenimine (PEPEI) as chelator for Cu2+ recovery in PEUF. The results show a maximum retention percentage over 97 % obtained with feed loading ratio of 208 mg Cu per g. of PEPEI at pH 6 (Llanos et al., 2008). Labanda et al. (2009) compared the ethoxylated PEI (EPEI) with polyvinyl alcohol (PVA), polyacrylic acid-co-maleic acid (PACM) and PEI on the recovery of Cr3+ in PEUF. They discovered that among these four types of chelating agents, PACM with carboxylic groups formed relatively stable irreversible complex with Cr3+, whereas alcoholic PVA with hydroxyl groups did not chelate with Cr3+ at all. The amines of PEI and EPEI formed hydroxo-complexes with Cr3+ ions under varying pH (Labanda et al., 2009). Since water-soluble polymers like PEI form highly stable metal complex in aqueous solution, it is not easy to be dissociated in low pH medium to achieve objective of metal recovery. A type of ‘intelligent’ polymeric microgel with thermo-responsive property known as poly(N-isopropylacrylamide-co-acrylic acid) (PNIPAM-co-AA) polymer hydrogel is applied as a novel chelating agent in PEUF for removal of divalent copper ions (Cu2+) from aqueous solution. This polymer hydrogel consists of a cross-linked network structure with a shape of globule conformation which makes each hydrogel structure a discrete nano-size gel particle in aqueous phase. It undergoes a temperatureinduced volume phase transition (VPT) process manifested as swelling-shrinking behaviour under varying temperature (Yang et al., 2004, Burmistrova et al., 2011, Chen et al., 2013, Zhang and Wang, 2009, Yamashita et al., 2003).

molecules which leads to expansion of the hydrogel network. As temperature increases above its VPTT, the hydrophilic groups are hidden and forming strong inner hydrogen bonds within hydrogel interior. The formation of short-range hydrogen bonds among ligands causes expulsion of water from hydrogel and shrinkage of the entire hydrogel network conformation. On the other hand, hydrophobic groups like isopropyl and methyl groups are gradually exposed to the exterior of hydrogel at temperature above VPTT, and the shrinking mechanism causes hydrogel to be separated from aqueous phase (Chen et al., 2013, Zhang and Wang, 2009, Yamashita et al., 2003). The thermo-responsive behaviour of PNIPAM-coAA polymer hydrogel before and after copper complexation is characterized by using dynamic light scattering (DLS) technique. The change of hydrodynamic diameter, Dh is plotted within temperature range of 25 – 50 °C. Figure 1(a) and (b) show the temperature-induced Dh curves for PNIPAM before and after complexing with copper ions, respectively. The Dh of hydrogel before complexation with copper decreases from 610 to 380 nm with increasing temperature from 25 to 50 ºC. After copper-hydrogel complexation, the sharp reduction in Dh above VPTT had disappeared. This indicates that the binding of Cu2+ with functional groups had replaced the water-associated hydrogen bonds. The ordinary trend of Dh change with temperature as in Figure 1(a) is disturbed. The curve in Figure 1(b) shows that the rate of Dh change is lowered for both heating and cooling cycles. More pronounced hysteresis effect is observed between heating and cooling curves in Figure 1(b). This is because copper chelation had reduced water-associated ligands and chelation of Cu2+ ion by different ligands causes certain extent of irreversible shrinkage (Chen et al., 2013, Zhang and Wang, 2009, Yamashita et al., 2003, Zhang et al., 2005, Sun et al., 2011). 4. FACTORS THAT AFFECT COMPLEXATION OF COPPER BY PNIPAM-co-AA

3. THERMO-RESPONSIVENESS OF PNIPAMco-AA POLYMER HYDROGEL

The affinity of chelating ligands in PNIPAM-co-AA toward copper ions in aqueous solution are sensitively influenced by several factors including pH, density of ligands (chelating functional groups) and temperature (thermo-responsive polymers) (Flora and Pachauri, 2010, Hancock and Martell, 1989).

Physically, the hydrogel swells below its volume phase transition temperature (VPTT) (~32 °C) because its hydrophilic functional ligands like amides (-CONH) and carboxylic (-COOH) and carbonyl (C=O) tend to form hydrogen bonds with water

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(a)

(b)

Fig. 1: Thermo-responsiveness of hydrogels: (a) before and (b) after copper complexation

observed from 26.6% to 53.4% when the solution pH changed from 4.0 to 5.0. It is because increase in pH value leads to increase in OH- hydroxyl ions from the base added (NaOH) that consume H+ ions in the solution. Consequently, lesser free H+ ions compete with Cu2+ ions for the chelation sites of ligands. From pH 6 to pH 9, copper removal was increased linearly from 63.2% to 87.6% due to the fact that hydroxides of copper started to form at pH 6 and it was increasing at higher pH. The dominating species of copper at pH 3 – 5 is Cu2+ while at pH 6 and above, insoluble Cu(OH)2 is dominating. At pH 6, Cu2+ is still dominant in the solution, but the presence of monohydroxo Cu(OH)+ which possess single charge was adsorbed to a greater extent than divalent Cu2+ and slight Cu(OH)2 were also formed under this condition. Thus, the increase in removal efficiency at pH 5 is owed to deprotonation of AA ligands, whereas at pH 6 and above, Cu(OH)+ and Cu(OH)2 species had significant effect towards enhancement of copper removal (Yao et al., 2010, Powell et al., 2007, 2011).

4.1. Effect of pH pH is a critical parameter for influencing the rate, efficiency and stability of metal-polymer complex formation by altering the charges of ligands. Most of the polymers bearing functional groups like –NH (amine) and -COOH (carboxylic acid) as chelating ligands are inactive for chelating metal ions at low pH (< pKa) due to protonation of the functional groups and electron donors are not available for donating electrons to metal ion. As pH increases above pKa of the acidic ligand of polymer, the functional group (ligand) is deprotonated and increases affinity of the ligand toward metal ion. Nevertheless, at high pH (≥ pH 6), the formation of insoluble metal hydroxide precipitates reduces free metal ions from being chelated by polymer ligands leading to a decline in metal-chelating efficiency of the polymer. In the case of copper, it exists as a hydrated tetracopper cationic complex, [Cu(H2O)4]2+ at pH ≤ 6 in aqueous solution. At pH range of 3 – 5, mononuclear complexes of one ligand (LM+) or two ligands (L2M) are formed. In condition where the concentration ratio of ligands to copper cations is low (