Fluid Waste Disposal (Environmental Science, Engineering and Technology)

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Fluid Waste Disposal Kay W. Canton (Editor) 2010. ISBN: 978-1-60741-915-0

ENVIRONMENTAL SCIENCE, ENGINEERING AND TECHNOLOGY SERIES

FLUID WASTE DISPOSAL

KAY W. CANTON EDITOR

Nova Science Publishers, Inc. New York

Copyright © 2010 by Nova Science Publishers, Inc. All rights reserved. No part of this book may be reproduced, stored in a retrieval system or transmitted in any form or by any means: electronic, electrostatic, magnetic, tape, mechanical photocopying, recording or otherwise without the written permission of the Publisher. For permission to use material from this book please contact us: Telephone 631-231-7269; Fax 631-231-8175 Web Site: http://www.novapublishers.com NOTICE TO THE READER The Publisher has taken reasonable care in the preparation of this book, but makes no expressed or implied warranty of any kind and assumes no responsibility for any errors or omissions. No liability is assumed for incidental or consequential damages in connection with or arising out of information contained in this book. The Publisher shall not be liable for any special, consequential, or exemplary damages resulting, in whole or in part, from the readers‘ use of, or reliance upon, this material. Independent verification should be sought for any data, advice or recommendations contained in this book. In addition, no responsibility is assumed by the publisher for any injury and/or damage to persons or property arising from any methods, products, instructions, ideas or otherwise contained in this publication. This publication is designed to provide accurate and authoritative information with regard to the subject matter covered herein. It is sold with the clear understanding that the Publisher is not engaged in rendering legal or any other professional services. If legal or any other expert assistance is required, the services of a competent person should be sought. FROM A DECLARATION OF PARTICIPANTS JOINTLY ADOPTED BY A COMMITTEE OF THE AMERICAN BAR ASSOCIATION AND A COMMITTEE OF PUBLISHERS. LIBRARY OF CONGRESS CATALOGING-IN-PUBLICATION DATA Fluid waste disposal / editor, Kay W. Canton. p. cm. Includes index. ISBN 978-1-61122-590-7 (eBook) 1. Sewage disposal. I. Canton, Kay W. TD741.F55 2009 628.3--dc22

Published by Nova Science Publishers, Inc.  New York

2009037445

CONTENTS Preface Chapter 1

Chapter 2

Chapter 3

Chapter 4

Chapter 5

Chapter 6

Chapter 7

Chapter 8

ix Treatment of Wastewater by Electrocoagulation Method and the Effect of Low Cost Supporting Electrolytes Lazare Etiégni, K. Senelwa, B. K. Balozi, K. Ofosu-Asiedu, A. Yitambé, D. O. Oricho and B. O. Orori

1

Application of Sulphate-Reducing Bacteria in Biological Treatment Wastewaters Dorota Wolicka

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Utilization of Water and Wastewater Sludge for Production of Lightweight-Stabilized Ceramsite Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren

83

Modelling and Observation of Produced Formation Water (PFW) at Sea D. Cianelli, L. Manfra, E. Zambianchi, C. Maggi and A. M. Cicero Disposal of Sulfur Dioxide Generated in Industries Using Eco-Friendly Biotechnological Process – A Review A. Gangagni Rao and P.N. Sarma Novel Biological Nitrogen-Removal Processes: Applications and Perspectives J.L. Campos, J.R. Vázquez-Padín, M. Figueroa, C. Fajardo, A. Mosquera-Corral and R. Méndez Application of Microbial Melanoidin-Decomposing Activity (MDA) for Treatment of Molasses Wastewater Suntud Sirianuntapiboon and Sadahiro Ohmomo Wastewaters from Olive Oil Industry: Characterization and Treatment L. Nieto Martínez, Gassan Hodaifa,M. Eugenia Martínez and Sebastián Sánchez

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viii Chapter 9

Chapter 10

Chapter 11

Contents Usability of Boron Doped Diamond Electrodes in the Field of Waste Water Treatment and Tap Water Disinfection Hannes Menapace, Stefan Weiß, Markus Fellerer, Martin Treschnitzer and Josef Adam Utilization of Biosolids as Fertilization Agents on Agricultural Land: Do the Obvious Benefits of Recycling Organic Matter and Nutrients Outweigh the Potential Risks? Veronica Arthurson Integrated Approach for Domestic Wastewater Treatment in Decentralized Sectors Rani Devi and R. P. Dahiya

Chapter 12

Biodegradation Characteristics of Wastewaters Fatos Germirli Babuna and Derin Orhon

Chapter 13

Batch Treatment of a Coffee Factory Effluent for Colour Removal Using a Combination of Electro-Coagulation and Different Supporting Electrolytes L. Etiégni, D. O. Oricho, K. Senelwa B. O. Orori, B. K. Balozi, K. Ofosu-Asiedu and A. Yitambé

Chapter 14

Water as a Scarce Resource: Potential for Future Conflicts M. A. Babu

Chapter 15

Recycling Wastewater After Hemodialysis: An Environmental and Cost Benefits Analysis for Alternative Water Sources in Arid Regions Faissal Tarrass, Meryem Benjelloun and Omar Benjelloun

Chapter 16

Chapter 17

Index

Pb (II) Ions Removal by Dried Rhizopus Oligosporus Biomass Produced from Food Processing Wastewater H. Duygu Ozsoy and J. Hans van Leeuwen Control of Plasticizers in Drinking Water, Effluents and Surface Waters Rosa Mosteo, Judith Sarasa, M. Peña Ormad and Jose Luis Ovelleiro

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249 265

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PREFACE Wastewater is any water that has been adversely affected in quality by anthropogenic influence. It comprises liquid waste discharged by domestic residences, commercial properties, industry, and/or agriculture and can encompass a wide range of potential contaminants and concentrations. In the most common usage, it refers to the municipal wastewater that contains a broad spectrum of contaminants resulting from the mixing of wastewaters from different sources. With the dwindling available water resources in the world coupled with high population growth, pressure is being exerted on water and wastewater plant managers the world over to find cost-effective methods to treat a wide range of wastewater pollutants in a diverse range of situations. This new and important book gathers the latest research from around the globe on fluid waste disposal with a focus on such topics as: wastewaters from the olive industry, application of sulphate-reducing bacteria in biological treatment wastewaters, electrocoagulation treatment method, usability of boron doped diamond electrodes in wastewater treatment and others. Chapter 1 - Coagulation and flocculation are traditional methods of treating of polluted water. Electrocoagulation (EC) presents a robust novel and innovative alternative in which a sacrificial metal anode doses water electrochemically. This has the major advantage of providing active cations required for coagulation, without necessarily increasing the salinity of the water. Electrocoagulation is a complex process with a multitude of mechanisms operating synergistically to remove pollutants from water. A wide variety of opinions exist in the literature for key mechanisms and reactor configurations. A lack of a systematic approach has resulted in a myriad of designs for electrocoagulation reactors without due consideration of the complexity of the system. A systematic, holistic approach is required to understand electrocoagulation and its controlling parameters (pH, temperature, conductivity, current density). This will enable a priori prediction of the treatment of various pollutant types. Electrocoagulation involves applying a current across electrodes in water. This results in the dissolution of the anode (either aluminum or iron). These ions then form hydroxides which complex with and/or absorb contaminants and precipitate out. The precipitate with the contaminants can then be removed from the water by settling and decantation or filtration. EC has the potential to be applied in many other areas besides the textile and semiconductor industry. It has been successfully tested in the pulp and paper industry, as well as tea and coffee processing. However over electrical potential within electrodes during electrocoagulation normally causes extra voltage, which wastes energy. There have been attempts to reduce this extra voltage which, in these days of World energy crisis, will render

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the electrocoagulation process uneconomical. The inclusion of supporting electrolyte such as NaCl achieves this. One of the methods pioneered by researchers at Moi University in Kenya is the use of wood ash leachate as supporting electrolyte which in some cases could reduce energy consumption by as much as 80%. Other supporting electrolytes tested are ash from bagasse and from coffee husks. These supporting electrolytes are relatively inexpensive, but they all generally result in large amount of coagulated sludge. Other supporting electrolytes such phosphate rock are less effective than wood ash, but they yield almost 50% less sludge after electrocoagulation. Most of the supporting electrolytes have an added advantage of reducing other wastewater pollution parameters such as BOD, COD, TSS, TS, turbidity, pH and color. Because of the inherent benefits of these low cost supporting electrolytes, electrochemical methods could be a credible alternative to more traditional wastewater treatment approaches. Chapter 3 - Disposal of wastewater treatment sludge (WWTS) and drinking-water treatment sludge (DWTS) is one of the most important environmental issues nowadays. Traditional options for WWTS and DWTS management, such as landfilling, incineration, etc., are no longer acceptable because they can cause many environmental problems. Conversion of WWTS and DWTS into useful resources or materials is of great interest and must be intensely investigated. To attain this goal, WWTS and DWTS were used as components for making ceramsite. Part I: SiO2 and Al2O3 were the major acidic oxides in WWTS and DWTS, so their effect on characteristics of ceramsite was investigated. Results show that WWTS and DWTS can be utilized as resources for producing ceramsite with optimal contents of SiO2 and Al2O3 ranging from 14–26% and 22.5–45%, respectively. Bloating and crystallization in ceramsite above 900 ℃ are caused by the oxidation and volatilization of inorganic substances. Higher strength ceramsite with less Na-Ca feldspars and amorphous silica and more densified surfaces can be obtained at 18%≤Al2O3≤26% and 30%≤SiO2≤45%. Part II: Fe2O3 and CaO were the major basic oxides, so their effect on characteristics of ceramsite was also investigated. The optimal contents of Fe2O3 and CaO are in the range of 5%–8% and 2.75%–7%, respectively. Higher strength ceramsite with more complex crystalline phases and fewer pores can be obtained at 6%≤Fe2O3≤8%. Lower strength ceramsite with more pores and amorphous phases can be obtained at 5%≤CaO≤7%, which implies that excessive Ca2+ exceeds the needed ions for producing electrical neutrality of silicate networks. Part III: To investigate stabilization of heavy metals in ceramsite, leaching tests were conducted to find out the effect of sintering temperature, pH, and oxidative condition. Results show that sintering exhibits good binding capacity for Cd, Cr, Cu, and Pb and leaching contents of heavy metals will not change above 1000 ℃. Main compounds of heavy metals are crocoite, chrome oxide, cadmium silicate, and copper oxide, which prove that stronger chemical bonds are formed between these heavy metals and the components. Leaching contents of heavy metals decrease as pH increases and increase as H2O2 concentration increases. Leaching results indicate that even subjected to rigorous leaching conditions, the crystalline structures still exhibit good chemically binding capacity for heavy metals and it is environmentally safe to use ceramsite in civil and construction fields. It is concluded from the 3 parts that utilization of WWTS and DWTS can produce high performance ceramsite, in accordance with the concept of sustainable development. Chapter 4 - Through the last decades, the ever increasing energetic demands have been accomplished by exploiting new natural reservoirs, including offshore oil and gas deposits located in marine coastal areas. During the extraction and production phases, large amounts of

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water are brought to the surface along with the hydrocarbons. These waters include the ‗formation water‘, that lies underneath the hydrocarbon layer, and ‗additional water‘ usually injected into the reservoirs to help force the oil to the surface. Both formation and injected waters, named ―produced formation waters‖ (PFWs), are separated from the hydrocarbons onboard offshore platforms and then disposed into the marine environment through ocean diffusers. PFWs contain several contaminants and represent one of the main sources of marine environment pollution associated with oil and gas production. This makes the study of PFW fate of paramount importance for a proper management of environmental resources as well as for planning and optimizing the discharge and monitoring procedures. In the first part of this chapter we provide a detailed description of the chemical characteristics of PFWs and their potential toxic effects and review the mixing processes governing their dispersion in the marine environment. In the second part of the work we briefly review past efforts in observing and modelling PFW spreading in the ocean. Finally, we propose a multidisciplinary approach, integrating in situ observations and numerical modelling, to assess dispersion of PFWs in space and time. As a case study we will refer to the results of a previous study conducted in the Northern Adriatic Sea, a sub-basin of the Mediterranean Sea, where a number of offshore natural gas (CH4) extraction platforms are currently active. Chapter 5 - Sulfur dioxide (SO2) is a known pollutant and responsible for various ill effects on living and non-living organisms. SO2 emissions can be reduced by using nonconventional energy sources or using conventional fuels containing less sulfur. However, under the present circumstances SO2 emissions cannot be completely avoided due to the reasons of rapid industrialization. Various technologies are available for the removal of SO2 from flue and waste gases. Most of these technologies fall under the category of physical, chemical or thermal. All these technologies generate secondary pollutants ending up in disposal problems and also cost prohibitive. Biotechnology offers relatively cheaper solutions for the conventional problems. Due to this reason, biotechnology is making in roads into the conventional treatment processes in all the fields. Over the last decade, efforts have been made to develop biotechnological alternatives to conventional physico- chemical processes for the removal of SO2 from flue gases known as Biological flue gas desulphurization (BIOFGD).SO2 from flue gas can be absorbed in a suitable organic media. In the aqueous phase SO2 would be converted to sulfite and some part may again be converted to sulfate due to the presence of dissolved oxygen. Therefore, the aqueous phase will be having both sulfate and sulfite, which can be reduced to sulfide using Sulfate Reducing Bacteria (SRB) under anaerobic conditions. The sulfide formed in the anaerobic reactor could be converted to elemental sulfur using Sulfur Oxidizing Bacteria (SOB) under partial microbial aerobic conditions. The elemental sulfur can be used either as a soil conditioner or raw material for industrial applications. Therefore, BIO-FGD process could be an environmentally benign and economically viable alternative for the disposal of SO2 emitted from the industries especially from power plants and refineries. The present article reviews the state of art of BIO-FGD process. Chapter 6 - Since the requirement for nutrient removal is becoming increasingly stringent, a high efficiency of nitrogen removal is necessary to achieve a low total nitrogen concentration in the effluent. Biological nitrification and denitrification processes are generally employed to remove nitrogen from wastewater. Unfortunately, these processes are

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not suitable to treat wastewater with a low COD/N ratio because it involves the addition of an external organic carbon source and, therefore, an increase of the operational costs. Several alternative processes for nitrogen removal can be applied in order to reduce partially (―nitrite route‖) or totally (anammox, autotrophic denitrification) the organic matter required. Such processes suppose not only an economical way to treat these wastewaters but they are also more environmentally friendly technologies (lower production of CO2, N2O and sludge; lower energy consumption). Up to now, they were basically applied to the return sludge line of municipal wastewater treatment plants (WWTPs). However, these processes could even be implemented in the actual WWTPs in order to achieve more compact and energy efficient systems. Their potential advantages can make them also feasible technologies to treat polluted ground water or to remove nitrogen compounds from recirculating aquaculture systems. Chapter 7 - This review will discuss the melanoidin-decomposing activity (MDA) among microorganisms. The focus will be on the potential use of the microbial-MDA to treat the wastewater discharged from factories using molasses as the raw material (molasses wastewater: MWW) because molasses is one of the most useful raw materials in various types of industries, such as the fermentation and animal feed industries. However, the wastewater discharged from factories using molasses contains a large amount of dark brown pigment, melanoidin pigment: MP, which is poorly decomposed and/or decolorized by normal biological treatment processes, such as the activated sludge or anaerobic treatment systems (anaerobic pond or anaerobic contact digester), because, the microorganisms in those wastewater treatment systems showed very poor MDA. The distribution of MDA among microorganisms and the mechanism of decomposing activities, in particular, were reviewed. Also, the application of the isolated strains having the MDA to treat molasses wastewater in the wastewater treatment plant was tested. Chapter 8 - Countries in the Mediterranean basin are among the main producers of olive oil. The elaboration of olive-oil is typically carried out by small companies in small facilities. The olive-oil plants produce high and variable amounts of residual waters of olives and oliveoil washing (OMW) that has a great impact in the environment. According to the procedure used different types of OMW with different chemical oxygen demand can be generated, the OMW from the three phase process (COD = 150 g O2 L-1) and the OMW from olives washing (COD = 0.8-4.5 g O2 L-1) and olive oil washing (COD = 1.1- 6 g O2 L-1) in the two-phase process. The uncontrolled disposal of OMW is a serious environmental problem, due to its high organic load, and because of its high content of microbial growth-inhibiting compounds, such as phenolic compounds. The improper disposal of OMW to the environment or to domestic wastewater treatment plants is prohibited due to its toxicity to microorganisms, and also because of its potential threat to surface and groundwater. These waters normally are stored in great rafts of accumulation for their evaporation during the summer. This solution among others until the moment dose not represent a definitive solution for this problem, especially as the administrations more and more demanding the preparation of this spill and the constructive quality of the rafts. Today, effective technologies have been proposed such as the chemical oxidation process using ferric chloride catalyst for the activation of H2O2 as a treatment of OMW produced from two-phase process. In the previous works the authors have described the experimental results on laboratory-scale. These results have been taken to pilotindustrial scale, making the chemical oxidation in the optimum conditions of operations: [H2O2] = 5% (w/v), using a ferric chloride catalyst with a relation of [FeCl3]/[H2O2] = 0.25

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(w/w), at OMW pH and environmental temperature. The final average value of COD obtained next to 370 mg L-1 (%CODremoval = 86.2%), and the water obtained can be destined to irrigation or disposed directly to the municipal wastewater system for their tertiary treatment. OMW from three-phase process does not allow direct chemical and biological purification for its content in phenolic compounds and generally used natural and forced evaporation process. Another way of using is the application of OMW nutrients to the growth of microorganisms such as microalgae. Chapter 9 - Over the past few years one main focus on the research efforts at the Institute for Sustainable Waste Management and Technology (IAE) has been on possible applications for reactors with boron doped diamond electrodes (BDD) in the field of (waste) water treatment. This article deals with the technical construction of the electrodes used (continuous reactor with a different number of plate electrodes), which were produced by a spin-off of the institute. The electrodes consist of conductible industrial diamond particles (< 250 µm), which are mechanically implanted on a fluoride plastic substrate. These electrodes showed a high mechanical and chemical stability in different test runs. At the institute, treatment methods for micro pollutants (e.g. pharmaceuticals and complexing agents) were developed with electrochemical oxidation by BDD. In this case test runs were made on laboratory scale and technical scale treatment units and elimination rates up to 99 % were achieved. In this project the analytic is partly provided by the ―Umweltbundesamt GmbH‖ (UBA), one of the project partners. This agency has been a project partner in different studies about pharmaceuticals in the ecosystem. These techniques could also be used for the waste water treatment of alpine cabins. Pilot projects have been set up. On the basis of these results a follow-up project was launched last October, in which an alternative treatment process for oilin-water emulsions and mixtures was developed by the usage of electrochemical oxidation with BDD. A third possible application is the disinfection of drinking water from contaminated ground and spring water. In this process oxidation agents like ozone or OH radicals produced in situ by the BDD reactor from the treated water are used to eliminate bacterial contaminants (for example e. coli) in the water. Chapter 10 - Treatment of wastewater, commonly performed at municipal sewage plants, generates sanitized water and sewage sludge. Anaerobic degradation of sewage sludge results in the production of different gases, including the economically valuable methane, and digested residue (biosolids) with potential value as a crop fertilizer. Traditionally, digested sewage sludge is disposed either into water, onto or into the earth or into the air. However, alternative exploitation of digested sewage sludge in agriculture has several advantages over commercial fertilizers, including environmental aspects benefiting agricultural sustainability and increased crop yield. Additionally, residue utilization is nearly always a cheaper option than disposal costs. Biosolids obtained from the treatment of municipal sewage sludge consist of a mixture of organic and mineral compounds that significantly affect soil microbial communities and their biogeochemical activities when applied as a crop fertilizer. The microorganisms influence soil quality through nutrient cycling, decomposition of organic matter and maintenance of soil structure, in turn, affecting agricultural and environmental quality, and subsequently, plant and animal health. Moreover, both soil and residue normally contain considerable quantities of microorganisms, including both beneficial and potentially human pathogenic species that may be supported by the new conditions in the soil. Thus, soil amended with biosolids may

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present a modified microbial community composition after some time and, hence, a modified ecosystem function. At the end of the present chapter, we discuss whether the potential risks of recycling biosolids to agricultural cropland are acceptable for consumers, producers and scientific expertise, in view of the resulting alterations in soil microbial diversity, activity and accompanying functions. Furthermore, optimal ways of managing the recycling process to achieve the most favourable balance of benefits and risks for the community are highlighted. Chapter 11 - The purpose of the present study was to design an integrated wastewater treatment system for a nalla (riverlet) flowing through Indian Institute of Technology Delhi (IITD), India, besides its cost estimation and comparison with the conventional wastewater treatment system. The design parameters for integrated aeration-cum-adsorption tank were worked out for 240 m3 / d flow rate of the wastewater. The important parameters used for the design included initial COD and BOD concentration in the influent, treatment time, adsorbent dose, pH, adsorbent particle size and the desired COD and BOD in the effluent after treatment as prescribed by Central Pollution Control Board, (CPCB) Delhi, India. All the design parameters of this system were similar to those of conventional system except for the replacement of aeration tank in conventional system by the aeration-cum-adsorption tank. The concentration of COD and BOD of the treated effluent by the integrated system were well within the permissible limits of CPCB standards (for COD it is 100 ppm and for BOD of 30 ppm) to discharge in the canal for irrigation purpose. It was worth mentioning here that the adsorbents used in the present study were based on discarded materials which were available free of cost. Of course, the cost of their transportation and processing should have been taken into account. The total cost estimated for the conventional system and the adsorption based system would be Rs. 198,312 and Rs. 141,275 respectively (including civil work, machinery, labour, adsorbent and miscellaneous). The cost difference for the two systems would be approximately Rs 57,037. This design of integrated system has resulted into saving of cost by 28 % over the conventional system. Thus, it is a good approach for saving of conventional energy in addition to saving the cost of treatment and can be applicable for any country for decentralized sector. Moreover, it is an open ended research and we can recommend more research by changing the adsorbents types and operating parameters to improve the model. Chapter 12 - The objective of this chapter is to put forward an overview of biodegradation characteristics of wastewaters by emphasizing the significance of COD fractionation. Recalcitrant COD fractions of effluents can be used as a tool to evaluate whether discharge standards can be met with a prescribed biological treatment. Moreover, the appropriate type of biological treatment applicable to the wastewater under investigation can be addressed and the performance of an existing biological treatment system can be appraised with reference to inert COD fractions. Besides recalcitrant COD fractions of segregated industrial effluent streams can be regarded as an essential input of a sound industrial wastewater management strategy adopting minimization at source philosophy. Last but not least, data on COD fractions can be used as a solid source of information for modelling studies that define the design and performance of biological treatment systems. In this context, COD fractionation data on a wide spectrum of activities ranging from various industrial sectors to hotels is presented. Segregated industrial wastewater streams together

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with domestic sewage and end-of-pipe industrial effluents are evaluated in terms of their biodegradation characteristics. Chapter 13 - In the present study, two types of colour removal systems were tested on effluent samples collected from a coffee pulping factory which discharged on average 15 m3 of wastewater daily with a colour index of about 2500 OH that was too high for direct discharge into a river in Kenya. The two colour removal systems used were: (i) electrolysis combined with wood ash or coffee husks leachate and (ii) electrolysis combined with phosphate rock solutions at a rate of 0.5 g/l to 4g/l. Phosphate rock is often used as agricultural liming agent. The surface area of the electrodes was set at close to 75 m2/m3 of effluent with a current density of 1,200 mA/m2. The experiments were laid out in a stratified random sampling design and the data were analysed using the Statistical Package for Social Scientists (SPSS) computer programme version 10.0. Electrolysis combined with phosphate rock (ELPHOS) proved to be the best process in terms of power consumption (68% reduction) compared with the 57% reduction by electrolysis combined with wood ash (ELCAS) and the 58% reduction by electrolysis combined with coffee husks ash (ELCHAS). Besides the 100% colour removal, ELPHOS also reduced other effluent physico-chemical parameters such as BOD, COD, TSS and TS by 79%, 80%, 69%, and 88% respectively. The analysis of ELPHOS treated wastewater showed that the mill could discharge an effluent that meets local discharge standards for colour requirements. It is recommended that recycling of the treated water by ELPHOS back to the factory for cleaning and washing purposes be considered since the quality meets the requirement for uses of fresh water for cleaning purposes. Furthermore, calculation of power consumption based on a scale-up batch reactor of 15 m3 proved less expensive to treat the factory effluent than a set of 12 one 100-L reactors similar to the one used in the field. Chapter 14 - The major aim of this paper is to review the major problems of water resources in the developing countries. It is based on problems related to population growth and pollution and how these are more likely to lead to future conflicts. We know that fresh water is only 3 % of the total global water and 78% of this is in glaciers. This makes it a scarce and precious resource which must be sustainably managed. The paper also analyses some of the already existing and potential conflicts based on water resources. It reviews the potential threats to Ugandan water resources and problems which are most likely to occur as a result of these threats. Factors hindering treatment of wastewater as a remedy to pollution in developing countries have also been discussed. The methodology used in this paper is based on literature review of the most current issues that affect water resources world-wide. The review is limited to scientific facts and no political factors affecting water resources have been included. It has been found that although Uganda is endowed with 66km2/year of renewable water resources, population increase, deforestation, degradation of wetlands and pollution are major threats to its water resources. Problems associated with water quality and quantities are more likely to result into internal conflicts which are bound to spread beyond Ugandan borders. Chapter 15 - Water is a vital aspect of hemodialysis. During the procedure, large volumes of water are used to prepare dialysate and to clean and reprocess machines. This paper evaluates the technical and economical feasibility of recycling hemodialysis wastewater for irrigation uses, such as watering gardens and landscape plantings. Water characteristics, possible recycling methods, and the production costs of treated water are discussed in terms of the quality of the generated wastewater. A cost-benefit analysis is also performed through

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comparison of intended cost with that of seawater desalination, which is widely used in irrigation. Chapter 16 - Heavy metal pollution is a serious problem in many developed and developing countries. Lead had been recognized as a particularly toxic metal and comes into water bodies mainly from metallurgical, battery, metal plating, mining and alloy industries. In order to minimize the impacts of this metal on human health, animals and the environment, lead-contaminated water and wastewater need to be treated before discharge to water bodies. This chapter concerns an investigation of potential usage of corn-processing wastewater as a new alternative low-cost substrate to produce biosorbent and evaluate this biosorbent to remove Pb(II) ions from aqueous solutions. For this aim, Rhizopus oligosporus cultivated on corn-processing wastewater and dried biomass of these fungi was used as an adsorbent. The adsorption experiments were conducted in a batch process and the effects of contact time (148 hours), initial pH (2-7), initial metal ion concentration (20-100 mg L-1) and adsorbent dosage (0.5-5 g L-1) on the adsorption were investigated. Pb (II) ion concentrations before and after adsorption were measured using Inductively Coupled Plasma-Mass Spectrometry. Maximum adsorption capacity was achieved at pH 5.0. The isothermal data of dried fungal biomass could be described well by the Langmuir equation and monolayer capacity had a mean value of 59.88 mg g-1. The pseudo-second order reaction model provided the best description of the data with a correlation coefficient 0.99 for different initial metal concentrations. This result indicates that chemical sorption might be the basic mechanism for this adsorption process and Fourier Transform Infrared Spectroscopy analyses showed that amide I and hydroxyl groups play an important role in binding Pb (II). Because of the high activation capacity of adsorbent and low cost of process dried R. oligosporus biomass presents a good potential as an alternative material for removal of Pb (II) ions from the aqueous solutions. Chapter 17 - The main objective of this research work is to determine the presence of di(2-ethylhexyl) phthalate, di(2-ethylhexyl) adipate and diisodecyl phthalate, in different water samples (drinking waters, effluents and surface waters). Different analytical methods were studied in order to know the best methodology for the quantification of these compounds. Solid-liquid and liquid-liquid extraction were investigated and finally the liquidliquid extraction and analysis by gas chromatography followed by mass spectroscopy was chosen because of offering the highest recovery rate. In the whole of this research study, the control of background pollution by reagents and material was extremely important. The problem of background pollution is more serious in the trace analysis of phthalates and adipates as a consequence of their presence in almost all equipment and reagents used in the laboratory. Respect to the control of the selected plasticizers in the different water samples, bis (2ethylhexyl) phthalate and bis (2-ethylhexyl) adipate were detected in drinking water, effluents and surface waters. On the other hand, diisodecyl phthalate was not detected in any sample.

In: Fluid Waste Disposal Editor: Kay W. Canton, pp. 1-48

ISBN: 978-1-60741-915-0 © 2010 Nova Science Publishers, Inc.

Chapter 1

TREATMENT OF WASTEWATER BY ELECTROCOAGULATION METHOD AND THE EFFECT OF LOW COST SUPPORTING ELECTROLYTES Lazare Etiégni1*, K. Senelwa1, B. K. Balozi1, K. Ofosu-Asiedu2, A. Yitambé3, D. O. Oricho1, and B. O. Orori1 1

Moi University, Department of Forestry & Wood Science, P. O. Box 1125 Eldoret, Kenya. 2 J.I.C., Dept. of Chem. Eng. Box 10099, Jubail Industrial City-31961, Kingdom of Saudi Arabia. 3 Kenyatta University, Department of Public Health P.O Box 43844-00100 Nairobi, Kenya

ABSTRACT Coagulation and flocculation are traditional methods of treating of polluted water. Electrocoagulation (EC) presents a robust novel and innovative alternative in which a sacrificial metal anode doses water electrochemically. This has the major advantage of providing active cations required for coagulation, without necessarily increasing the salinity of the water. Electrocoagulation is a complex process with a multitude of mechanisms operating synergistically to remove pollutants from water. A wide variety of opinions exist in the literature for key mechanisms and reactor configurations. A lack of a systematic approach has resulted in a myriad of designs for electrocoagulation reactors without due consideration of the complexity of the system. A systematic, holistic approach is required to understand electrocoagulation and its controlling parameters (pH, temperature, conductivity, current density). This will enable a priori prediction of the treatment of various pollutant types. Electrocoagulation involves applying a current across electrodes in water. This results in the dissolution of the anode (either aluminum or iron). These ions then form hydroxides which complex with and/or absorb contaminants and precipitate out. The precipitate with the contaminants can then be *

Corresponding author: E-mail: [email protected]

2

Lazare Etiégni, K. Senelwa, B. K. Balozi et al. removed from the water by settling and decantation or filtration. EC has the potential to be applied in many other areas besides the textile and semiconductor industry. It has been successfully tested in the pulp and paper industry, as well as tea and coffee processing. However over electrical potential within electrodes during electrocoagulation normally causes extra voltage, which wastes energy. There have been attempts to reduce this extra voltage which, in these days of World energy crisis, will render the electrocoagulation process uneconomical. The inclusion of supporting electrolyte such as NaCl achieves this. One of the methods pioneered by researchers at Moi University in Kenya is the use of wood ash leachate as supporting electrolyte which in some cases could reduce energy consumption by as much as 80%. Other supporting electrolytes tested are ash from bagasse and from coffee husks. These supporting electrolytes are relatively inexpensive, but they all generally result in large amount of coagulated sludge. Other supporting electrolytes such phosphate rock are less effective than wood ash, but they yield almost 50% less sludge after electrocoagulation. Most of the supporting electrolytes have an added advantage of reducing other wastewater pollution parameters such as BOD, COD, TSS, TS, turbidity, pH and color. Because of the inherent benefits of these low cost supporting electrolytes, electro-chemical methods could be a credible alternative to more traditional wastewater treatment approaches.

INTRODUCTION With the dwindling availability of water resources in the World coupled with high population growth, pressure is being exerted on water and wastewater plant managers the world over to find cost-effective methods to treat a wide range of wastewater pollutants in a diverse range of situations. Traditionally coagulation, flocculation and lagooning have been used as chemical and biological processes with varying degrees of success to treat polluted waters. However a more cost-effective and proven method to clean an ever widening range of water pollutants, on-site, and with minimum additives, is required for sustainable water and wastewater management. Electrocoagulation treatment of water seems to fit this description. Colloidal dispersions in water or wastewater often referred to as sols consist of discrete particles held in suspension by their extreme small size (1-200 nm), state of hydration (chemical combination with water), and surface electric charge. The chemistry of coagulation and flocculation is primarily based on the electrical properties of the particles. Like charges repel each other while opposite charges attract. Particles finer than 0.1 µm (1x10-7 m) in water or wastewater remain continuously in motion due to electrostatic charges (often negative) which cause them to repel each other. There are two types of colloids - hydrophilic and hydrophobic. Hydrophilic are readily dispersed in water and their stability depends on the affinity for water rather than the slight negative charge they possess. Hydrophobic colloids on the other hand have no affinity for water and their stability depends on the charge they possess, usually positive. The electrostatic repulsion between the colloidal particles leads to a stable sol. The surface or primary charge of colloidal particles comes from charged groups within the particles or the adsorption of charged particles. The sign and magnitude of the surface charge depends on the character of colloids, the pH (the lower the pH the more positive the charge becomes), the ionic strength and the characteristics of the water or wastewater. The surface of the colloid has a certain δ-potential (zeta potential) which is the magnitude of the charge at the surface of shear. The δ-potential is derived from the diffused double-layer

Treatment of Wastewater by Electrocoagulation Method…

3

theory applied to hydrophobic colloids (Figure 1), and can be estimated using Smoluchowski‘s (1872-1917) electrokinetic mobility equation:

δε μ=

--------------------------------- (1) (1)

ε where μ = the electrophoretic mobility δ = zeta potential ε = the electric permittivity ε = the viscosity of the water or wastewater

Zeta potential can also be calculated using the following relationship for electrostatic force

4πqd δ=

--------------------------- (2) D

(2)

q = charge per unit area d = thickness of the layer surrounding the shear surface through which the charge is effective π = pi (= 3.142857) D = dielectric constant of the liquid + + + + +

+

+

+ +

Stern layer

+ +

+

+

+ +

Surface shear

+

+

+ +

+

+ +

Particle + +

+

+

+

Bulk of solution n

+ + + +

+

+

+

+ +

+ + +

+ +

Zeta potential +

Fixed layer of ions

Figure 1. Diffused double layer.

Electric potential surrounding particle Diffusion layer of counterions

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Lazare Etiégni, K. Senelwa, B. K. Balozi et al.

The diffused double layer (Figure 1) consists of two parts: an inner region, also referred to as Stern layer, which includes ions bound relatively strongly to the surface (including specifically adsorbed ions) and an outer, diffuse or movable region in which the ion distribution is determined by a balance of electrostatic forces and random thermal motion. The potential in this region, therefore, decays as the distance from the surface increases until, at sufficient distance, it reaches the bulk solution value, conventionally taken to be zero. The repulsive force of the charged double layer scatters particles thus preventing agglomeration. Particles with high zeta potential have a very stable sol. The zeta potential is the overall charge a particle acquires in a specific medium. In other words, it is a measure of the magnitude of electrical charge surrounding the colloidal particles. The magnitude of the zeta potential gives an indication of the potential stability of the colloidal system. Zeta potential can be equated to the amount of repulsive force which keeps the particles in suspension. If the zeta potential is large, then more coagulants will be needed to destabilize colloidal particles. If all the particles have a large negative or positive zeta potential they will repel each other and there is dispersion stability. When particles have low zeta potential values, there is no force to prevent the particles coming together and there is dispersion instability. A dividing line between stable and unsable aqueous dispersions is generally taken at either +30 or -30mV.

FACTORS AFFECTING ZETA POTENTIAL There are several factors that can affect zeta potential

1. pH In aqueous media, the pH of a sample is one of the most important factors that affect its zeta potential. A zeta potential value on its own without defining the solution conditions is virtually meaningless. A zeta potential versus pH curve will be higher or positive at low pH and lower or negative at high pH. There may be a point where the plot passes through zero zeta potential. This point is called the isoelectric point and is very important from a practical consideration. It is normally the point where the colloidal system is least stable.

2. Conductivity The thickness of the double layer (κ-1) depends upon the concentration of ions in solution and can be calculated from the ionic strength of the medium. The higher the ionic strength, the more compressed the double layer becomes. The valence of the ions will also influence double layer thickness. A trivalent ion such as Al3+ will compress the double layer to a greater extent in comparison to a monovalent ion such as Na+. Inorganic ions can interact with charged surfaces in one of two distinct ways (i) non-specific ion adsorption where they have no effect on the isoelectric point and (ii) specific ion adsorption, which will lead to a

Treatment of Wastewater by Electrocoagulation Method…

5

change in the value of the isoelectric point. The specific adsorption of ions onto a particle surface, even at low concentrations, can have a dramatic effect on the zeta potential of the particle dispersion. In some cases, specific ion adsorption can lead to charge reversal of the surface.

3. Concentration of a Formulation Component The effect of the concentration of a formulation component on the zeta potential can give information to assist in formulating a product to give maximum stability. The influence of known contaminants on the zeta potential of a sample can be a powerful tool in formulating the product to resist flocculation for example.

COAGULATION Schulze, in 1882, first showed that colloidal systems could be destabilized by the addition of ions having a charge opposite to that of the colloid (Benefield et al., 1982). Coagulation in water or wastewater chemistry is a process in which a chemical referred to as a coagulant is added to destabilize dispersed colloidal particles so that they agglomerate. Coagulation experiments using natural products such as Moringa oleifera have also been tried with varying degrees of success (Kasser et al., 1990; Ogutveren et al., 1994; Ndabigengesere et al., 1995; Mohammed, 2001; Bhuptawat and Chaudhari, 2003). The objectives of coagulation are to (i) destabilize suspended and colloidal particles to enhance their removal through sedimentation and filtration and (ii) to precipitate dissolved maters i.e. PO43-, color, natural organic matter (NOM). Coagulation process may require several reaction steps: (i) hydrolysis of multivalent metal ions; (ii) adsorption of hydrolysis species at the solid-solution interface for the destabilization of colloidal particle (reduction of zeta potential); (iii) aggregation of destabilized particles by interparticle bridging; (iv) aggregation of destabilized particles by particle transport and van der Waals‘ forces; (v) ―aging‖ of flocs formed in the process; and (vi) precipitation of metal hydroxides (Stumm and O‘Melia, 1968).

ELECTROCOAGULATION Electrocoagulation is a process that applies a current across electrodes through a liquid, using a variety of anode and cathode geometries, including plates, balls fluidized bed spheres, wire mesh, plates (either aluminum or iron), rods, and tubes. This results in the dissolution of the anode (Equation 3 & 12). These ions then form hydroxides which complex with and/or absorb contaminants and precipitate from water or wastewater. They are subsequently removed by surface complexation and electrostatic attraction according to the following equations:

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Lazare Etiégni, K. Senelwa, B. K. Balozi et al.

WITH IRON ELECTRODES In acidic medium, Fe(s) → Fe2+(aq) + 2e-

(Equation 3)

4Fe2+(aq) + 10 H2O(l) + O2 (g) → 4 Fe (OH)3(s) + 8H+ (aq)

(Equation 4)

Cathode:

2H+ (aq) + 2e- → H2 (g)

(Equation 5)

Overall:

4 Fe(s) + 10 H2O(l) + O2 (g) → 4Fe(OH)2(s) + 4H2 (g)

(Equation 6)

Anode:

In alkaline medium, Anode:

Fe(s)

Fe2+ (aq) + 2e-

Fe2+(aq) + 2OH-(aq) 2H2O(l) + 2e-

Cathode: Overall:

Fe (s) + 2H2O(l) Floc + H2 (g)

(Equation 7)

Fe(OH)2(s)

(Equation 8)

H2(g) + 2OH-(aq)

(Equation 9)

Fe(OH)2(s) + H2(g) Floats

(Equation 10) (Equation 11)

WITH ALUMINUM ELECTRODES In acidic medium Anode:

Al (s) 2Al3+ (aq) + 4H2O(l) + O2 (g)

Cathode:

2H+ (aq) + 2eOverall: 2Al(s) + 4H2O(l) + O2 (g)

Al3+ + 3e-

(Equation 12)

2Al(OH)3(s) + 2H+ (aq) (Equation 13) H2 (g)

(Equation 14)

2Al(OH)3(s) + 2H2(g)(Equation 15)

In alkaline medium Anode:

Al (s) Al3+ (aq) + 3OH- (aq)

Cathode: Overall:

2H2O(l) + 2e2Al(s) + 3H2O(l)

Al3+ + 3e-

(Equation 16)

Al(OH)3(s)

(Equation 17)

H2(g) + 2OH-(aq)

(Equation 18)

Al(OH)3 + 3/2H2 (g)

(Equation 19)

Treatment of Wastewater by Electrocoagulation Method…

7

The cation hydrolyses in water to form a hydroxide. The following equations (20 to 23) are an illustration of this phenomenon in the case of aluminum: pH

Al3+ + H2O AlOH2+ + H2O Al(OH)2+ + H2O Al(OH)30 + H2O

AlOH2+ + H+ Al(OH)2+ + H+ Al(OH)30 + H+ Al(OH)4- + H+

(Equation 20) (Equation 21) (Equation 22) (Equation 23)

ELECTROCOAGULATION MECHANISMS The electrocoagulation overall mechanism is a combination of mechanisms that operate concurrently or in series but synergistically. The main mechanism may vary throughout the dynamic process as the reaction progresses, and will almost certainly shift with changes in operating and environmental parameters and pollutant types. Highly charged cations destabilize any colloidal particles by the formation of polyvalent polyhydroxide complexes. These complexes have high adsorption properties, forming aggregates with pollutants. The pollutants presumably act as a ligand to bind with iron or aluminum ions resulting in the formation of amorphous polymeric complexes (hydroxo-complexes). These compounds with a large specific surface area are very active and able to coagulate and adsorb pollutants soon after their in situ generation (Rajeshwar and Ibanez 1997; Scott, 2001). Besides the generation of polyvalent cations described above, electrocoagulation includes also the production of electrolysis gases that are hydrogen and oxygen (Equation 5, 6, 9, 10, 14, 15 & 19). Evolution of hydrogen gas aids in mixing and flocculation. Once the floc is generated, the electrolytic gas binds to and creates a buoyant force on the floc leading to its flotation and ultimately to the removal of the pollutant as a floc - foam layer at the liquid surface (Equation 11). Other flocs that are heavier settle at the bottom of the reactor. There are many ways in which species can interact in solution: 1. Migration to an oppositely charged electrode (electrophoresis) and aggregation due to charge neutralization. 2. The cation or hydroxyl ion (OH-) forms a precipitate with the pollutant. 3. The metallic cation interacts with OH- to form a hydroxide, which has high adsorption properties thus bonding to the pollutant (bridge coagulation). 4. The hydroxides form larger lattice-like structures that sweep through the water (sweep coagulation). 5. Oxidation of pollutants to less toxic species. 6. Removal by electroflotation and adhesion to bubbles (Figure 2). Electrocoagulation process has been around for some time. The process was proposed before the turn of the last century with Vik et al. (1984) describing a treatment plant in London built in 1889 (for the treatment of sewage by mixing with seawater and electrolyzing). In 1909, Harries (cited in Vik et al., 1984) in the United States, received a patent for wastewater treatment by electrolysis with sacrificial aluminum and iron anodes. Matteson and Dobson (1995) described a device of the 1940‘s, the ―Electronic Coagulator‖

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Lazare Etiégni, K. Senelwa, B. K. Balozi et al.

which electrochemically dissolved aluminum (from the anode) into solution, reacting this with the hydroxyl ion (from the cathode) to form aluminum hydroxide. The hydroxide flocculates and coagulates the suspended solids purifying the water. A similar process was used in Britain in 1956 for which iron electrodes were used to treat river water (Matteson and Dobson, 1995). Because of its capability to remove several types of water pollutants, the recent thirty years have seen an explosion of journal article reports on electrocoagulation methods probably due to new and more stringent environmental regulations on a wide range of water and wastewater pollutants. This has further translated into a number of electrocoagulation devices, designed to purify water or wastewater, being put on the market.

REACTIONS WITHIN THE ELECTROCOAGULATION REACTOR Several distinct electrochemical reactions are produced independently within the electrocoagulation reactor. They are as follows: Emulsion breaking, resulting from the oxygen and hydrogen ions that bond into the water receptor sites creating a water insoluble complex that separate water from pollutants. Seeding, resulting from the anode reduction of metal ions that become new centers for larger, stable, insoluble complexes that precipitate as complex metal ions. Bleaching by the oxygen ions produced in the reaction chamber oxidizing pollutants such as dyes, cyanides, biohazards, chlorolignins from pulp and paper mill effluent. DC power supply

Coagulation & Flocculation

Cathode

Anode

Flocs

H2

Sediments Figure 2. Electrocoagulation process interactions (Hydrogen discharge at the cathode generates gas micro-bubbles that cause the floatation of flocs and the increase of pH).

Treatment of Wastewater by Electrocoagulation Method…

9

Electron flooding of the water that eliminates the polar effect of the water complex, allowing colloidal materials to precipitate. The increase of electrons creates an osmotic pressure that ruptures bacteria, cysts, and viruses. Oxidation reduction reactions that are forced to their natural end point within the reactor which speeds up the natural process. Electrocoagulation induced pH swings toward neutral although this will not always be the case and will depend on the type of electrolyte used.

TYPE OF ELECTRODES Electrode material can subtancially affect the performance of an electrocoagulation reactor. The heart of EC is the dimensionally stable oxygen evolution anode which is usually expensive. The anode material determines the cation introduced into solution. Several researchers have studied the choice of electrode material with a variety of theories as to the preference of a particular material. The most common electrodes were aluminum or iron plates as described by Vik et al. (1984) and Novikova and Shkorbatova (1982). Do and Chen (1994) have compared the performance of iron and aluminum electrodes for removing color from dye-containing solutions. Their conclusion was that the optimal electrocoagulation conditions varied with the choice of iron or aluminum electrodes, which in turn was determined by initial pollutant concentration and pollutant type.

STIRRING RATE Bazrafshan et al. (2008), while comparing chromium removal efficiency with iron and aluminum electrodes, showed that removal efficiency of chromium with aluminum electrodes was lower than chromium removal efficiency with iron electrodes. Metal consumption equally was much lower with aluminum than with iron electrodes. Conversely, power consumption was lower with aluminum than with iron electrodes for the same concentration of pollutant. However, as the chromium concentration in the solution increased to 500.0 mg/L , the consumption of the electrode reduced, but efficient chromium removal occurred due to the large amount of flocs formation that helped sweep away chromium. For example, iron electrode consumption for the initial concentration of 5.0 mg/l and voltage of 40 V was 9.01 g while for an initial concentration of 500.0 mg/L it was 7.70 g (Bazrafshan et al., 2008). The highest efficiency of chromium removal (for both iron and aluminum electrodes) was measured in acidic medium (pH = 3) for an initial chromium concentration of 500.0 mg/L and at lower concentrations, the removal efficiency was almost complete at all pH values. At high chromium concentration, however, the complete removal would have required longer time i.e. higher power consumption. Some researchers have investigated the relationship between ―size‖ of the cation introduced and removal efficiency of organic waste (Baklan and Kolesnikova, 1996; Vlyssides et al., 1997). The size of the cation produced (10-30μm for Fe3+ compared to 0.05-1 μm for Al3+) was suggested to contribute to the higher efficiency of iron electrodes. Their conclusion was based on a single experiment, however, using chemical absorption of oxygen

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Lazare Etiégni, K. Senelwa, B. K. Balozi et al.

as the only measure. Hulser et al. (1996) observed that electrocoagulation is strongly enhanced at aluminum surfaces in comparison to steel. This is attributed to a higher efficiency due to the in situ formation of dispersed aluminum-hydroxide complexes through hydrolysis of the aluminate ion, which does not occur with steel electrodes. Tsai et al. (1997) employed Fe and Al anodes to simultaneously utilize electrocoagulation, responsible for removal of high molecules, and oxidation during treatment of a raw leachate. Iron anodes provided better COD removal at low applied voltages than did aluminum (Englehardt et al., 2006). As a general statement the efficiency of aluminum or iron electrodes will depend on the specific type of pollutant and also on the different set of operating parameters (Kobya et al., 2003).

ELECTRODE PASSIVATION One of the greatest operational issues with electrocoagulation is electrode passivation. The passivation of electrodes is of concern for the longevity of the process. Passivation of aluminum electrodes has been widely reported in the literature (Nikolaev et al., 1982; Osipenko and Pogorelyi, 1977). The latter also observed that during electrocoagulation with iron electrodes, deposits of calcium carbonate and magnesium hydroxide were formed at the cathode and an oxide layer was formed at the anode. Nikolaev et al. (1982) investigated various methods of preventing electrode passivation and suggested the following options for its control: Changing polarity of the electrode; Hydromechanical cleaning; Introducing inhibiting agents; Mechanical cleaning of the electrodes. According to these researchers, the most efficient and reliable method of electrode maintenance was to periodically mechanically clean the electrodes or wash the electrodes with 8% sulfuric acid between runs in batch which for large-scale, continuous processes is a challenging issue. Corrosion promoters such as Cl - ions have been found to induce thinning of passive layer, enhance dissolution and promote depassivation (spontaneous depassivation). Other types of electrodes with a wide range of materials have been tested for electrocoagulation process. These materials include: Graphite, Platinum oxide, Iridium oxide, lead oxide, tin oxide, boron doped diamond (BDD). Graphite electrodes are deemed to be cheap but unstable and for most part ineffective (Barisa et al., 2009). They become easily fouled during the electrocoagulation process and this reduces their effectiveness. Platinum and Iridium oxide electrodes are too expensive and ineffective. Electrodes made of lead oxide (PbO 2) and tin oxide (SnO 2) are easy to manufacture but they are highly unstable. Boron doped diamonds are materials suitable for use as anodes in the electrocoagulation of organic compounds. Due to their very high resistance to deactivation via fouling and extreme electrochemical stability they show no significant corrosion even under high current densities. They have good chemical, mechanical and

Treatment of Wastewater by Electrocoagulation Method…

11

thermal resistance and a wide electrochemical potential window in aqueous solutions. Above all they can provide very high current efficiencies. Diamond coated electrodes have been investigated worldwide over the past number of years with notable results (Fryda et al., 2003). It is possible to vary electrical properties of diamond from semiconductor (very wide band gap) to close to metallically conductive by varying the boron doping level (1019-1021 cm-3). The most important electrochemical properties of BDD electrodes are their very high corrosion stability in electrochemical applications and their extremely high overvoltage for water electrolysis (Fryda et al., 1999). This large working potential window in aqueous electrolytes provides the possibility of producing strong oxidizing solutions with extremely high efficiency. As reported by Michaud and Comninellis (2000), compared to other electrode materials, BDD electrodes produce hydroxyl radicals on their surface with higher current efficiency. These hydroxyl radicals completely mineralize organic impurities in water or wastewater, such as oil, cooling fluid, toxic compounds (Tennakone et al., 1995). As diamond electrodes are both stable as anodes and cathodes, it is possible to reverse polarity in order to prevent calcium build-up on the electrode surface. Through the use of diamond electrodes, it is possible to obtain an electrochemical process which, without the addition of further chemicals, results in an environmentally friendly and relatively maintenance-free method for the treatment of waste water. Nonetheless, despite the promising results with respect to effectiveness and energy efficiency which have been demonstrated for wastewater treatment, electrosynthesis and electroplating, BDD electrodes remain extremely expensive. A new anode coated IrOx−Sb2O5−SnO2 onto titanium has also been proposed (Xueming et al., 2002). Accelerated life test showed that the electrochemical stability of the Ti/IrOx−Sb2O5−SnO2 anode containing only 2.5 mol % of IrO x nominally in the activated coating was even higher than that of the conventional Ti/IrO x anode. Its service life for electroflotation application is predicted to be about 20 years. Voltametric investigation demonstrated that the Ti/IrO x−Sb2O5−SnO2 anode could provide fast electron transfer. The present anode had a fork-like design and arranged in an interlocking manner with the cathode with a similar shape. Such an innovation in electrode configuration and arrangement is claimed to allow bubbles produced at both electrodes to be dispersed into wastewater flow quickly and, therefore, enhances the effective contact between bubbles and particles, favorable for high flotation efficiency. In addition, the novel electrode system reduces the interelectrode gap to 2 mm, a spacing that is technically difficult for a conventional electrode system (Xueming et al., 2002). This small gap results in a significant energy saving. Easy maintenance is another advantage of this novel electrode system.

AREAS OF APPLICATION OF ELECTROCOAGULATION According to Can et al. (2006), electrocoagulation has been proposed in recent years as an effective method to treat various wastewaters such as: landfill leachate, restaurant wastewater, saline wastewater, tar sand, paper mill effluent, coffee factory effluent, tea factory effluent, oil shale wastewater, urban wastewater, laundry wastewater, nitrate and

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Lazare Etiégni, K. Senelwa, B. K. Balozi et al.

arsenic bearing wastewater, and chemical mechanical polishing wastewater. The electrocoagulation process can successfully remove a wide range of pollutants in a much shorter time than conventional treatment methods (Ogutveren et al., 1994; Kongsricharoern and Polprasert, 1995, ). They include: removal of metals, oil, BOD, TSS, TDS, FOG, color etc., from wastewater before final disposal, thus reducing or eliminating discharge surcharges; reconditioning antifreeze by removing oil, dirt, and metals; reconditioning brine chiller water by removing bacteria and fat; pretreatment before membrane technologies such as reverse osmosis, ultrafiltration and nanofiltration; preconditioning boiler makeup water by removing silica, hardness, TSS; reconditioning boiler blow down by removing dissolved solids eliminating the need for boiler chemical treatment; recycling water, allowing closed loop systems; harvesting protein, fat and fiber from food processor waste streams; de-watering sewage sludge and stabilizing heavy metals in sewage, lowering freight and allowing sludge to be land applied; conditioning and polishing drinking water; removing chlorine and bacteria before water discharge or reuse (Cenkin and Belevstev ,1985; Biswas and Lazarescu,1991; Browning,1996; Adhoum et al., 2004).

COLOR Color is found mostly in surface waters, although some groundwater inside deep wells may also contain color that is noticeable (APHA-AWWA, 1992; AWWA, 1999). Many domestic and industrial wastewaters are rarely colorless and the color levels depend on the industrial process and the age of the wastewater i.e. the travel time in the collection and treatment system (Kim et al., 2005). The pulping and bleaching of wood for example generally produce large amounts of wastewaters that contain lignin derivatives and other dissolved wood by-products. Lignin derivatives which are usually brownish in color remain resistant to biological degradation during wastewater treatment. The brownish color of a pulp and paper mill effluent is mainly attributed to products of lignin polymerization formed during pulping and bleaching operations. These chromophoric groups are mainly quinonic types with conjugated double bonds originating from pulping processes (Luner et al., 1970). When disposed of into natural watercourses, they add color which persists for great distance. Additionally, colored effluents from pulp and paper mills for example result in reduced photosynthetic activity, increased long term BOD, increased water treatment cost for users downstream, and increased toxicity (Springer et al., 1995). Several studies have been carried out to determine the effectiveness of EC in color removal. In general, the findings indicate that EC is more cost effective than normal or conventional coagulation. Moreover, other wastewater pollution parameters are reduced (Orori et al., 2005; Kashefialasl et al., 2006, Oricho et al., 2008). Electrocoagulation combined with wood ash or bagasse ash has also been applied on tea factory effluent. In one study by Maghanga (2008) on tea factory effluent, the treated effluent COD, BOD and electrical conductivity were reduced by 96.6%, 42.4%, and 20.9% respectively. Supporting electrolytes from wood ash, phosphate rock and bagasse ash further reduced power consumption by between 64% and 16%, confirming the effectiveness of this process.

Treatment of Wastewater by Electrocoagulation Method…

13

ADVANTAGES OF ELECTROCOAGULATION (EC) Electrocoagulation has several advantages that are as follows: EC produces effluent with less total dissolved solids (TDS) content compared to chemical treatments. If this water is reused, the low TDS level contributes to a lower water recovery cost. EC requires simple equipment and is easy to operate with sufficient operational latitude to handle most problems encountered during its running. Wastewater treated by EC can give palatable, clear, colorless and odorless water. Sludge formed by EC tends to be readily settable and easy to de-water, because it is composed of mainly metallic oxides/hydroxides. Flocs formed by EC are similar to chemical floc, except that EC floc tends to be much larger, contains less bound water, is acid-resistant and more stable, and therefore, can be separated faster by filtration. The EC process can remove the smallest colloidal particles, because the applied electric field sets them in faster motion, thereby facilitating their agglomeration and subsequent coagulation. The EC process often avoids uses of chemicals and so there may be no problem of neutralizing excess chemicals and no possibility of secondary pollution caused by chemical substances added at high concentration as when chemical coagulation of wastewater is used alone. The gas bubbles produced during electrolysis can carry the pollutant to the top of the solution where it can be more easily concentrated, collected and removed. The electrolytic processes in the EC cell are controlled electrically and with no moving parts, thus requiring less maintenance.

DISADVANTAGES OF ELECTROCOAGULATION (EC) High capital cost has often been cited as one of the major disadvantages of EC although labour requirement may also be high when running an EC batch reactor. Higher voltages and thus high specific energy consumption are also seen as a big disadvantage of the system. The final deficiency of this process relates to the fact that an EC reactor is an electrochemical cell whose performance is directly related to the operational state of its electrodes. As mentioned earlier, they vary widely in design and mode of operation- from simple vertical plate arrangements to packed-bed style reactors containing various metallic packings, and in material used (Ogutveren et al., 1992; Barkley et al., 1993). Potential for electrode passivation, thus slow reaction rates is another draw back because passivation impedes dissolution which normally provides the coagulants in situ. Electrode passivation, specifically of aluminum electrodes, has been widely observed and acknowledged as detrimental to reactor performance (Osipenko and Pogorelyi, 1977; Novikova and Shkorbatova, 1982). This formation of an inhibiting layer, usually an oxide, on the electrode surface prevents metal dissolution and electron transfer, thereby limiting coagulant addition to the solution. Over time, the thickness of this layer increases, reducing the efficacy of the electrocoagulation

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Lazare Etiégni, K. Senelwa, B. K. Balozi et al.

process as a whole. The use of new materials, different electrode types and arrangements (Pretorius et al., 1991; Mameri et al., 1998) and more sophisticated reactor operational strategies (such as periodic polarity reversal of the electrodes mentioned above) have led to significant reductions in the impact of passivation. The issue, however, is still seen as a serious potential limitation for applications where a low-cost, low maintenance water treatment facility is required.

DESIGN OF ELECTROCOAGULATION UNITS The inherent complexity of the electrocoagulation reactions makes it difficult to model and control this process. Adequate scale-up parameters, a systematic approach to the optimization and a priori prediction for the performance of the electrocoagulation reactor are yet to be established. A literature survey reveals that previously each ―new‖ system has been considered separately on an individual basis. The key driver for the development of any particular application of this process has generally been the removal of a specific pollutant i.e. color, heavy metal, COD, tannin etc. There has been little or no attempt to provide a holistic approach to electrocoagulation. Consequently, despite more than a century‘s worth of applications, many of them deemed successful, the science and engineering behind EC reactor design is still largely empirical and heuristic. It has failed to take full advantage of the potential success and incorporation in the understanding behind the science of electrocoagulation. A literature review indicates that EC reactors can be configured as batch or continuous and that the majority of reactors reported so far fall in the latter category with continuous feed and outflow operating under pseudo-steady state. Electrocoagulation systems require amperage to treat the water. The amount of amperage drawn is dependent upon the conductivity of the water or wastewater. If the water is not conductive then no amperage will be used. The system should be designed with adequate wiring and electrical capacity to deliver adequate amperage if needed by a particular water stream.

PHYSICAL DESIGN ISSUES There has been a range of laboratory, pilot and industrial scale electrocoagulation units produced. The designs range from fully integrated units to ‗stand alone‘ reactors. The electrocoagulation process has been combined with many units including microfiltration, dissolved air flotation (DAF), sand filtration and electroflotation. Obviously, pre- and postwater treatment impacts significantly on the performance of the electrocoagulation reactor. The design of the electrocoagulation process influences its operation and efficiency (Holt et al., 2005). The design phase should consider the following physical factors: Continuous versus batch operation Reactor geometry Reactor scale-up Current density

Treatment of Wastewater by Electrocoagulation Method…

15

GEOMETRY Geometry of the reactor affects operational parameters including bubble path, flotation effectiveness, floc formation, fluid flow regime and mixing/settling characteristics. From the literature, the most common approach involves plate electrodes (aluminum or iron) and continuous operation. Water is dosed with dissolved metal ions as it passes through the electrocoagulation cell. A downstream unit is often required to separate pollutant and water.

SCALE-UP ISSUES One of the cornerstones of chemical engineering is to establish key scale-up parameters to define the relationships between laboratory and full-scale equipment. The surface area to volume ratio (S/V) is a significant scale-up parameter. Electrode area influences current density, position and rate of cation dosing, as well as bubble production and bubble path length. Mameri et al. (1998) reported that as the S/V ratio increases the optimal current density decreases. However, the S/V ratio was not widely reported. Some of the values reported are listed in Table 1 below: The values reported here seem empirical with no specific criteria for their choice. A more rigorous and consistent approach is clearly required to establish a set of design characteristics for Electrocoagulation reactors. The prime differentiator between pollutant removal by settling or flotation would seem to be the current density employed in the reactor. A low current produces a low bubble density, leading to a low upward momentum flux—conditions that encourage sedimentation over flotation (Holt et al., 2002). As the current is increased, so does the bubble density resulting in a greater upwards momentum flux and thus more likely removal by flotation. Other researchers such as Zolotukhin (1989) scaled up an electrocoagulation-flotation system from laboratory to industrial scale. The following dimensionless scale-up parameters have been chosen to ensure correct sizing and proportioning of the reactors: Reynolds number – indication of the fluid flow regime; Froude number – indication of buoyancy; Weber criteria – indication of the surface tension; Gas saturation similarity; Geometric similarity. Table 1. S/V values reported in the literature. Reference (Author) Amosov et al. Osipenko and Pogorelyi Novikova and Shkorbatova Orori et al. Oricho et al. Maghanga

Year 1976 1977 1982 2005 2008 2008

S/V (m2/m3) 30.8 18.8 42.5 80.0 75.5 18.2

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Lazare Etiégni, K. Senelwa, B. K. Balozi et al.

Horizontal Flow

Vertical flow

Figure 3. Types of electrodes set-up during EC.

Electrodes during EC can be set up as parallel vertical or horizontal sheets as can be seen in Figure 3. The turbulence generated by the gases at the anode and cathode can be used in both types of flow. However, vertical flow allows for more improved separation by electroflotation as compared with horizontal flow.

FACTORS AFFECTING ELECTROCOAGULATION PROCESSES Several studies have shown that electrocoagulation is quite complex and may be affected by several operating parameters such as pollutant concentrations, initial pH, electrical potential voltage, COD, turbidity, pollutant type and concentration, bubble size and position, floc stability and agglomerate, size and type of supporting electrolyte. The complexity and number of possible interactions are highlighted in Figure 2.

EFFECT OF PH ON ELECTROCOAGULATION Optimal pH reported for electrocoagulation reactions varies significantly. These discrepancies probably derive from the complex and variability of wastewater composition, and the different operating conditions used in the electrocoagulation studies. It has been established that pH has a considerable effect on the efficiency of the electrocoagulation process (Springer et al., 1995, Chen et al., 2000, Li et al. 2001). The wastewater pH determines the speciation of metal ions and influences the state of other species in solution and the solubility of products formed. The pH of the medium also changes during electrocoagulation process, as observed by other investigators. This change depends on the type of electrode material and initial pH and alkalinity. In a study by Bazrafshan et al. (2008) on the removal of Chromium VI from synthetic chromium solutions by electrocoagulation

Treatment of Wastewater by Electrocoagulation Method…

17

using aluminum electrode, it was observed that there was an increased in the solution pH for an initial pH of less than 7. The increase was ascribed to hydrogen evolution at the cathodes contrary to Chen et al. (2000) assertion that the pH increase is due to the release of CO2 from wastewater owing to H2 bubble disturbance. At low pH, wastewater is over saturated with CO2 which can be released during H2 evolution, causing a pH increase. In addition, if the initial pH is acidic, reactions would shift towards a pH increase (Bazrafshan et al. 2008). During the same experiment, in alkaline medium (pH > 8), the final pH did not vary considerably but a slight drop was recorded. This result concurs with previous published works and suggests that electrocoagulation can act as a pH buffer (Gao et al.,2005). In the same study of chromium removal by electrocoagulation carried out over a wide range of Cr concentrations, it was also observed that the influent pH did not significantly affect the removal efficiencies of Cr VI. This means that for practical applications, pH adjustment before treatment is not required. In another study by Springer et al. (1995) on the effect of pH on the color removal reaction by electrocoagulation, it was found that higher pH slowed the electrocoagulation reaction, thereby increasing power consumption. In a separate study on color removal from a pulp and paper mill effluent, Orori (2003) found that decreasing the original effluent pH led to a significant reduction in power consumption during electrocoagulation combined with wood ash leachate. Lowering pH from 12.0 to pH 10.0 significantly (P 0.05) reduced power consumption by between 20 to 21% during electrochemical removal of a paper mill effluent color. It was postulated that decreasing the original effluent pH increased ionisation of wastewater, which increased the rate of iron (II) ions production at the anode and hydrogen at the cathode. Consequently, decreasing pH led to increased production of positively charged iron (II) ions, which attracted the negatively colored flocci (Springer et al., 1995). Thus increased production rate of these ions led to an increase in the rate of color removal at lower pH than at higher pH. Therefore lower pH facilitated color removal and lowered electrical power consumption. Li et al. (2001) reported that COD removal was at least 20% higher at pH 4.0 than at pH 8.0 after a 4-hour electrolysis. Vlyssides et al. (2003) found that pH was the most significant operational parameter in electrolyzing leachate, compared to Clconcentration, temperature, applied voltage, SO42- concentration and leachate input rate. Lower pH favored COD removal and saved energy consumption within the range pH 5.5 – 7.5. The disagreement in these investigations suggests further work, perhaps in terms of the mechanisms by which pH affects COD removal in leachate electrolysis. Theoretically, it can be stated that acidic conditions decrease the concentrations of CO32and HCO3- , both well-known scavengers of OH radical generated at the anodes (Li et al., 2001), while alkaline conditions promote the Cl-→Cl2→ClO-→Cl- redox cycle. Therefore, low pH may enhance direct oxidation, while high pH may enhance indirect oxidation (Wang et al., 2001). Thus, solution pH influences the overall efficiency and effectiveness of electrocoagulation. An optimal pH seems to exist for a given pollutant, with optimal pH values ranging from 6.5 to 7.5 (Holt et al.,2002). Kashefialasl et al. (2006) showed that the maximum efficiency of color removal during the treatment of dye solution containing colored index acid yellow 36 by electrocoagulation using iron electrodes was observed at pH range 7–9 as expected considering the nature of the reaction between ferrous and hydroxide ions. When the pH of solution was lower than 6, Fe(OH)3 was in soluble form (Fe+3) and when it was higher than 9, Fe(OH)3 was in soluble form {Fe(OH)4-} and because Fe(OH)3 played a major role in removing color, when pH of

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Lazare Etiégni, K. Senelwa, B. K. Balozi et al.

solution was 8, color removal was the highest. The dye solution with different initial concentrations in the range of 20-60 mg/l was treated by EC at optimized current density and time. In contrast, other investigators have found that pH variation does not considerably alter COD removal in leachate electrolysis. Chiang et al. (1995a) have reported that the pH effect on chlorine/hypochlorite production efficiency was insignificant over the range pH 4-10 during an electrolysis experiment in saline water conducted to help elucidate the mechanism of electrolyzing leachate. Cossu et al. (1998) found that a pseudo-first-order rate constant for COD reduction in real leachate increased only slightly at pH 3, compared with pH 8.3. Also, Wang et al. (2001) have reported that at pH 8.9 and 10, COD removal was approximately 4% higher than at pH 7.5, not a very significant effect.

EFFECT OF CURRENT DENSITY ON ELECTROCOAGULATION Current density (i) is the current delivered to the electrode divided by the active area of the electrode. Varying the current easily controls this parameter. Current density determines both the rate of electrochemical metal dosing to the water and the electrolytic bubble density production. Current densities ranging from 10 to 2000 A/m2 have been reported (Holt et al., 2005). The majority of the sources used for this write-up report a current density in the range 10 – 150 A/m2. Different current densities are desirable in different situations. Current densities reported for electrochemical oxidation of leachate ranged from 5 to 540 mA/cm2 (Englehardt et al., 2006). It is reported that at least 5 mA/cm2 is required to achieve effective oxidation of organics in leachate. Table 2. Percent of chromium removal during electrocoagulation process using aluminum electrodes (Initial concentration = 50 mg l−1). T = 60 min 98.62 98.74 98.88 98.40 98.44 98.72 92.00 97.64 98.34

T = 40 min 94.78 95.64 95.80 89.76 91.72 95.72 90.80 92.18 92.58

Source: Bazrafshan et al. 2008, with permission

T = 20 min 83.76 85.64 88.98 82.76 83.14 83.46 64.60 77.00 81.80

Voltage (V) 20 30 40 20 30 40 20 30 40

pH 3

7

10

High current densities are desirable for separation processes involving flotation cells or large settling tanks, while small current densities are appropriate for electro-coagulators that are integrated with conventional sand and coal filters. A systematic analysis will be required to define and refine the relationship between current density and desired separation effects. Current density (current per unit area of electrode) in an electrochemical process indicates gross reaction rate. For example under weaker oxidative conditions, leachate may darken and

Treatment of Wastewater by Electrocoagulation Method…

19

brown precipitates may form at the anode (Cossu et al., 1998; Li et al., 2001). Increasing current density improves COD and NH3-N treatment efficiencies at the same charge loading. Bazrafshan et al. (2008) showed that increasing electrocoagulation voltage increased the removal efficiency of Chromium, which was also helped by higher pHs as can be seen in Tables 1 and 2. Chiang et al. (1995b) reported that during electrolytic treatment of leachate, COD removal at 25 mA/cm2 was approximately 50% higher than that observed at 6.25 mA/cm2, for the same charge loading (1.178 x 105 Coulombs/L). This is probably due to the fact that increasing current density during electrolysis enhances chlorine generation, which may have been responsible for subsequent removal of pollutants (Costaz et al., 1983; Chiang et al., 1995a). Li et al., (2001) have shown that the effect of current density on treatment was not evident between 30 and 120 mA/cm2 at a low Cl- concentration (1650 mg/L), but became noticeable when Cl- concentration reached the 5000 mg/L level. Table 3. Percent of chromium removal during electrocoagulation process using aluminum Electrodes (Initial concentration = 500 mg l−1). T = 60 min 25.60 35.80 83.00 20.40 24.60 80.80 23.00 26.80 52.00

T = 40 min 22.0 27.00 71.20 19.60 20.40 64.60 13.80 22.00 41.20

Source: Bazrafshan et al. 2008, with permission.

T = 20 min 21.80 24.80 51.80 13.60 17.80 41.00 8.80 12.80 32.00

Voltage, (V) 20 30 40 20 30 40 20 30 40

pH 3

7

10

Figure 4. Effect of current density on the efficiency of color removal from a solution with concentration of the dye = 50 ppm (Source: Kashefialasl, et al., 2006 with permission).

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Lazare Etiégni, K. Senelwa, B. K. Balozi et al.

This result corroborates the importance of indirect oxidation during the electrolytic treatment of leachate. In addition, Moraes et al. (2005) reported that color removal from leachate strongly depended upon current density. Color removal efficiency at 116 mA/cm2 was five times higher than that at 13 mA/cm2, after 180 minutes of electrochemical treatment. In the treatment of dye solution containing colored index acid yellow 36 by electrocoagulation using iron electrodes Kashefialasl et al. (2006) showed that as current density increased so did color removal from the dye solution up to a certain maximum as shown in Figure 4. During electrocoagulation, electrical current not only determines the coagulant dosage rate but also the bubble production rate and size and the floc growth, which can influence the treatment efficiency by electrocoagulation (Letterman et al., 1999; Holt et al., 2002). This is ascribed to the fact that at higher voltage the amount of anode material oxidized increases, resulting in a greater amount of precipitate for the removal of pollutants. In addition, it has been demonstrated that bubble density increases and their size decreases with increasing current density resulting in a greater upwards flux and a faster removal of pollutants and sludge flotation (Khosla et al., 1991). As the current decreased, the time needed to achieve similar efficiencies increases. This expected behavior is explained by the fact that the treatment efficiency is mainly affected by charge loading (Q = It), as reported by Chen et al. (2000). However, the cost of the process is determined by the consumption of the sacrificial electrode and the electrical energy. It has also been established that for a given time, the removal efficiency increased significantly with increase of current density. The highest electrical potential normally produces the quickest treatment.

EFFECT OF THE CONCENTRATION OF POLLUTANTS Several investigations have shown that the initial concentration of pollutants has a bearing on the efficiency of the electrocoagulation process (Orori, 2003, Etiegni et al., 2007, Mahvi and Bazrafshan, 2007). A set of experiments was performed with different initial concentrations of chromium to determine the time required for its removal under various conditions of electrocoagulation process (Bazrafshan et al. 2008). SPP1

Power Consumption (MWh))

45

SPP2

40 SPP3

35

SPP4

30

SPP5

25 20 15 10 5 0 15

20

25

30

35

40

45

o

Temperature ( C)

Figure 5. Effect of temperature on power consumption by ELCAS at five sampling points along a pulp and paper Mill effluent treatment system (Source: Orori, 2003 with permission).

Treatment of Wastewater by Electrocoagulation Method…

21

The results obtained at different electrical potentials showed that initial concentration of chromium may have an effect on the efficiency of its removal and for higher concentration of chromium, higher electrical potential or more reaction time is needed. On the other hand, if the initial concentration increases, the time required should increase too. It is clear from Tables 1 & 2 that at higher concentrations, longer time is needed for removal of chromium, but higher initial concentrations of chromium were reduced significantly in relatively less time compared to lower concentrations. The time taken for its reduction thus increases with the increase in concentration. This can be explained by the theory of dilute solution. In dilute solution, formation of the diffusion layer at the vicinity of the electrode slows the reaction rate, but in concentrated solution the diffusion layer has no effect on the rate of diffusion or migration of metal ions to the electrode surface (Chaudhary et al., 2003). Chromium removal with respect to time by electrocoagulation process at different pH levels is shown in Tables 1 & 2.

EFFECT OF DISTANCE BETWEEN THE ELECTRODES

Metal removal (%)

Numerous research work have been conducted on the effect of electrode distance on the removal of wastewater contaminants by EC (Springer et al., 1995; Ecobar et al.,2006). For some researchers, the electrode gap did not seem to have an impact on the electrocoagulation process although at the narrowest of gap, the clogging of electrodes appeared to reduce the rate of reaction (Springer et al., 1995).

Distance between the electrodes (cm) Source: Escobar et al., 2006, with permission Figure 6. Effect of electrode gap on the removal of (A) Cu, (B) Pb and (C) Cd Current density=36 Am2, Electrolysis time =10 min, Conductivity = 900 mS/cm.

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Lazare Etiégni, K. Senelwa, B. K. Balozi et al.

EFFECT OF GAP BETWEEN ELECTRODES Escobar et al. (2006) while studying the optimization of the electrocoagulation process for the removal of copper, lead and cadmium in natural waters and simulated wastewater found that increasing the gap between the electrodes reduced metal removal due to a decrease in current flow and coagulant generation (Figure 6). An optimal distance of 2.0 cm for removal of lead and 2.5 cm for removal of cadmium and copper was identified for subsequent experiments. It will appear that each solution has an optimum electrode gap that should be determined experimentally before the EC can be optimized. In general, small gap results in a significant energy saving but makes it difficult to clean the electrode surface in conventional EC

EFFECT OF TURBIDITY In order to study the effect of turbidity (10, 50 and 200 NTU) on removal efficiency of cadmium a set of experiments was performed with different initial concentrations of cadmium (5, 50 and 500 mg l-1). The results obtained at optimum condition (pH=10, time reaction = 60 min and voltage = 40 V) showed that the removal efficiency for various concentrations of cadmium was fairly unchanged and hence electrocoagulation process can be applied efficiently for cadmium removal in the presence of turbidity (Mahvi and Bazrafshan, 2007).

EFFECT OF TEMPERATURE Raising temperature during electrocoagulation increases the rate of reaction (Shenz et al., 2006). Springer et al. (1995) showed that the time required for color removal reaction through electrocoagulation to reach 0.2 Absorbance Units (AU) was cut approximately by half by increasing temperatures from 23oC to 80oC (12 min vs 7 min). Orori (2003) studied the effect of temperature on color removal by electrocoagulation combined with wood ash leachate (ELCAS) and the results are shown in Figure 5. It was found that at 40oC color removal consumed less than 50% the electric power used at 20oC by ELCAS treatment. This was attributed to fast movement of electrons at higher temperatures compared to low temperatures.

EFFECT OF SUPPORTING ELECTROLYTES The underlying principle of EC (Figure 2) is the generation of cations by the dissolution of sacrificial anodes that induce flocculation of the dispersed pollutants contained by the zeta potential reduction system (Calvo et al., 2003; Mollah et al., 2004). During EC processes, high energy can be consumed leading to longer and slower reaction rates. For the EC to be effective, various types of electrodes and configurations have been tested. Several studies have also been conducted to determine the impact of certain additives such as supporting electrolytes during EC (Eichhorn et al.,1996). A supporting electrolyte (SE) is used to increase conductivity in the majority of all electroanalytical or electrosynthetic experiments

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23

such as electrocoagulation in aqueous and non-aqueous solutions (Lund et al., 1991; Fry, 1996). While a large number of experiments have been performed with electrodes under conditions where no SE was deliberately added, it is increasingly common practice to operate an EC in the presence of a certain amount of ions such as chloride or ammonium or salts such as NaCl, Na2SO4 and NaNO3 (Lopes et al., 2004; Orori et al., 2005; Shenz et al., 2006; Englehardt et al., 2006; Uğurlu, 2006; Oricho et al., 2008; Yildiz et al. 2008). Some of the electrolytes used in past experiments are shown in Table 4 and their respective effects on effluent color removal. Hu et al., (2003) carried out an experiment on defluoridation by EC and studied the effect of coexisting anions. The results showed that the type of dominant anion had a direct impact on the EC defluoridation reaction. Defluoridation efficiency was nearly 100% and most of the fluoride removal reaction occurred on the surface of the anode in the solution without the coexisting anions, due to the electro-condensation effect. In the solutions with co-existing anions, most of the defluoridation took place in the bulk solution. The residual fluoride concentration was a function of the total mass of Al3+ liberated. It was found that sulfate ions inhibited the localized corrosion of aluminum electrodes, leading to lower defluoridation process because of lower current efficiency. However the presence of chloride or nitrate ions prevented the inhibition of sulfate ions, and the chloride ions were more efficient. Different corrosion types occurred in different anion-containing solutions and the form of corrosion affected the kinetic over-potential of the EC reaction. When the concentration of NaCl salt or any other supporting electrolyte in solution increases, solution conductivity increases. Consequently, with respect to the solution voltage if any SE is added: V = EC - EA- δA- δC - IRcell - IRcircuit

(24)

where: EC EA δA δC IRcell IRcircuit

= Electrical potential difference at the cathode = Electrical potential difference at the anode = Zeta potential at the anode = Zeta potential at the cathode = voltage-drop across the cell = voltage-drop across the circuit

the necessary voltage for access to a certain current density will reduce, and the consumed electrical energy will be decreased (Kashefialasl et al., 2006). Excess SE affects the current in the bulk of the solution, which is maintained mostly by the ions of the SE, and migration effects on charged substrates can be neglected. The SE can also have some affects on the double layer reducing the Zeta potential of the substrate ions and helping their agglomeration or coagulation. Orori et al. (2005) showed that when the volume of wood ash leachate increased during color removal from a pulp and paper mill effluent, the power consumption reduced considerably by almost 80%. Similar results were also obtained by Etiégni et al. (2007) and Oricho et al. (2008).

24

Lazare Etiégni, K. Senelwa, B. K. Balozi et al. Table 4. Effect of the electrolyte concentration on the efficiency of color removal. Type of Electrolyte *NaCl *BaCl2 *KCl *NaBr *KI *Na2SO3 *Na2CO3 **Wood ash leachate ***Phosphate rock ****Bagasse ash leachate *****NH4NO3 *****Na2SO4 **Alum **Ca(OH)2

Applied voltage (v) 2.9 5.1 2.7 4.2 5.3 3.8 4.2 23 23 24 3.9 3.5 23 23

Conductivity (μS/cm) 19.13 9.67 18.75 11.77 9.18 13.8 15.8 4823.12 1150-1730 310

Color removal (%) 83 77 80 80 79 76 76 100 100 100

3456.23 3358.34

100 94 100 100

*Source: Kashefialasl et al., 2006: (Current density =127.8A/m2, Time of electrolyses =6min) **Source: Orori , 2003 (Current density= 250 A/m2, Time of electrolysis= 150 s) ***Source: Etiégni et al., 2007: (Current density = 222.2 A/m2, Time of electrolysis = 55 s) **** Source: Maghanga, 2008: (Current density = 55 A/m2, Time of electrolysis = 4 min) ***** Source: Lopes et al., 2004: (Current density = 2 mA/cm2, Time of electrolysis = 70-96 hrs)

30

Power (Whr)

25

OSA

NSA

OSC

NSC

OSR

NSR

20 15 10 5

0 0

1000

2000

3000

4000

5000

Electrolyte Dosage (g/m3 )

Source: Etiégni et al., 2009 Figure 7. Effect of electrolyte volume on the power consumption.

It appears that leachates from wood ash contain a wide range of ions or supporting electrolytes that may be helping or assisting the electrocoagulation reaction (Figure 7). Chou et al. (2009) studied the effect of NaCl concentration on the removal efficiency of indium

Treatment of Wastewater by Electrocoagulation Method…

25

(III). They showed that there was an increase removal efficiency up to 83% when NaCl (used as supporting electrolyte) concentration was 8 g/l. The concentration of supporting electrolyte was adjusted to the desired levels by adding a suitable amount of NaCl to the synthetic wastewater. Increasing the concentration of the supporting electrolyte from 0 to 200ppm led to an increase in indium (III) ion removal efficiency, whereas with the concentration of the supporting electrolyte increasing, the specific energy consumption decreased by almost 80%. When the concentration of the supporting electrolyte increased, the solution ohmic resistance decreased, so the current required to reach the optimum applied voltage diminished, decreasing the consumed energy (Chou et al. (2009). Although some SEs are available commercially, they can be extracted from material otherwise considered as waste. Several research papers have been recently published on the use of leachate from ash emanating from wood, coffee husk or bagasse as supporting electrolyte (Orori et al., 2005; Etiégni et al., 2007, Oricho et al., 2008).

OPERATING COST ANALYSIS OF ELECTROCOAGULATION PROCESSES Several studies have been carried out on the operating cost of electrochemically treated wastewater (Bayramoglu et al., 2004; Can et al., 2006; Bayramoglu et al., 2007). In a study by Bayramoglu et al. (2004) for the treatment of textile wastewater by EC using aluminum and iron electrode materials, the effect of wastewater characteristics and operational variables on the technical performances of COD and turbidity removal efficiencies as well as on the EC operating cost were determined.. Only direct costs such as material (electrodes and chemical reagents) and energy costs were considered for the calculation of the operating cost. Other cost items such as labor, maintenance and solid/liquid separation costs, depreciation of fixed investment such as rectifier and electro-coagulators were not taken into account. This simplified cost equation was used to evaluate the effect of various process variables on the operating cost: Operating cost + aCenergy + bCelectrode

(25)

where Cenergy and Celectrode, were consumption of energy and electrode material per kg of COD removed, which are normally obtained experimentally. Unit prices, a and b, determined for a specific market are as follows: a= electrical energy price and b= electrode material price for aluminum or for iron. Using equation 6, Bayramoglu et al. (2004) found that for iron electrode, the operating cost decreased initially with pH until pH = 5, where it remained constant up to pH=7, beyond which it increased (Figure 8). For aluminum electrodes, the EC cost increased with initial pH.

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Lazare Etiégni, K. Senelwa, B. K. Balozi et al.

Operating cost $/kg of COD

Figure 8 shows the effect of initial pH on the EC operating cost.

Initial pH Source: Bayramoglu et al. (2004) with permission Figure 8. Effect of initial pH on EC cost.

EFFECT OF CONNECTION MODES ON EC OPERATING COST When cells are set-up in an electrocoagulation process, one can choose from different modes or connections depending on the required voltage, the expected output and the overall efficiency of the EC system. Kobya et al. (2007) studied the effect of wastewater pH, current density and operating parameters for two sacrificial electrode materials, Fe and Al, and three electrode connection modes - namely monopolar-parallel (MP-P), monopolar-serial (MP-S) and bipolar-serial (BP-S) on the EC operating cost. The highest consumption of electrode material occurred with bipolar series mode (BP-S); approximately 0.27 kgm−3 for Fe electrode and between 0.18–0.23 kgm−3 for Al electrode. Monopolar parallel (MP-P) mode showed the lowest electrode consumption for both electrode Fe and Al materials: 0.12 kgm−3 for Al electrode and 0.16 kgm−3 for Fe electrode (Kobya et al., 2007). When the consumption of energy was compared for the three modes, as seen in Figure 10, only a minor change was observed with pH for all of the systems using Fe or Al electrodes. MP-S and BP-S modes exhibited high consumptions of energy because of the serial connection that required higher potential. When MP-P mode was used, it consumed the lowest energy or approximately 0.63 kWhm−3 and 0.7 kWhm−3 for Fe and Al electrode respectively. The effect of the initial pH on amount of sludge production is depicted in Figure 12. Sludge amounts vary from 0.65 to 1.0 kgm−3 for Fe electrode and from 0.9 to 1.3 kgm−3 for Al electrode (Kobya et al., 2007). In general, more sludge was produced with BP-S mode than with MP-P mode because of high electrode material consumed leading to high power consumption and higher electro-coagulant produced in situ. MP-P mode for both

Treatment of Wastewater by Electrocoagulation Method…

27

electrode materials was therefore economically more feasible owing to its low electrical energy consumptions and amount of sludge produced (Figure 12).

EFFECT OF ELECTRICAL CONDUCTIVITY OF EC COST Bayramoglu et al. (2004) studied the impact of electrical conductivity on the operating cost of a dye wastewater treatment system using two sets of electrodes (Figure 11). For both electrode materials, operating cost decreased with increasing conductivity and the decrease was almost similar for iron and aluminum electrodes with only a slight difference at 3500 μS/cm for aluminum electrode. For aluminum, the percentage of the electrode consumption cost with respect to the total cost was nearly constant as 76%. For iron, on the other hand, this ratio increases from 33 to 58%, with increasing conductivity from 1000 to 4000 μS /cm. The decrease in operating cost was probably due to a decrease in solution ohmic resistance. As SE increased, lower current required to reach the optimum applied voltage leading to the overall decreased of consumed energy (Chou et al. (2009).

Figure 9. Different types of electrode connection modes: a-Monopolar parallel (MP-P), b-Monopolar Serial (MP-S), c-Bipolar parallel (BP-P) modes. (Source : Kobya et al., 2007, with permission).

EFFECT OF RETENTION TIME ON EC COST Kobya et al. (2006) studied the effect of detention time during the treatment of potato chips manufacturing wastewater by electrocoagulation. They found that both energy and electrode consumption increased with retention time. Retention time is therefore likely to affect the operating cost of EC.

EFFECT OF CURRENT DENSITY ON EC COST In a study aimed at determining the impact of current density on operating cost of EC Kobya et al. (2007) showed that, in the case of iron, current density did not have a substantive impact on the performances of MP-P and MP-S modes; COD removal reached a maximum of

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Lazare Etiégni, K. Senelwa, B. K. Balozi et al.

67% with MP-P mode for a current density of 50Am−2. However, for aluminum electrode, the effect of the current density was more pronounced on COD removal, especially for MP-P mode and that lower current densities were more favourable. For example a 30Am−2 was preferred with MP-S mode.

Source: Kobya et al. (2003) with permission Figure 10. Effect of initial pH on energy consumption.

EFFECT OF POLYELECTROLYTE AND SE ON THE EC OPERATING COST As a general rule, EC operating cost has been found to reduce with the addition of polyelectrolyte up to an optimum concentration beyond which it usually rises, although this will also depend on the type of polyelectrolyte. Can et al. (2006) showed alum and polyaluminum chloride (PAC) increased operating EC operating cost when their concentration increased (Figure 13). However, Orori et al. (2005) found that increasing the concentration of wood ash leachate reduced power consumption and reduced operating cost, although the cost of electrode replacement and sludge removal was not included in the overall operating cost calculations.

Operating cost $/kg COD

Treatment of Wastewater by Electrocoagulation Method…

Conductivity, μS/cm Source: Bayramoglu et al. (2004) with permission. Figure 11. Effect of wastewater conductivity on EC cost.

Source: Bayramoglu et al. (2007) with permission

Figure 12. Effect of initial pH on sludge formation

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Lazare Etiégni, K. Senelwa, B. K. Balozi et al.

Operating cost ($/kg COD)

30

Polyelectrolyte dosage (kg/m3) Source: Can et al. (2006) with permission Figure 13. Effect of polyelectrolyte addition on EC operating cost.

WOOD ASH LEACHATE USED AS A SUPPORTING ELECTROLYTE What Is Wood Ash? Wood ash is the residue powder left after the combustion of wood. The main producers of wood ash are wood industries, power plants, homesteads especially in Third World countries. Large amount of this residue are produced every day. Typically 6-10 percent of burned wood results in ash. Wood ash is commonly disposed of in landfills or agricultural lands, but with rising disposal costs ecologically friendly alternatives are becoming more attractive. These alternatives will be based on the ash composition. It has been demonstrated that wood ash composition is a function of the wood combustion temperature as can be seen in Table 5 (Etiégni and Campbell, 1991). Wood combustion produces a highly alkaline ash that can be used to neutralize acidic effluent. As can be seen in Table 5 below Ca, K, Mg, Mn, Fe and Na are important elements found in wood ash. Misra et al., (1993) analyzed samples of wood ash using Inductively Coupled Plasma Emission Spectrometer (ICPES) and X-ray diffraction (XRD) to identify the minerals present in wood ash. A list of the compounds identified in ash is shown below in Table 6. The low temperature ash at 600oC showed strong peaks corresponding to calcium carbonate. Pine and aspen ash contained relatively higher amounts of potassium compared to poplar ash and showed strong peaks corresponding to K2Ca(CO3)2. Pine ash contained calcium manganese oxide, Aspen ash had sulfates of calcium and potassium, and poplar ash, silicates of K, Mg, and Ca. At higher temperatures (1000oC) where most industrial wood-fired boilers operate, with the dissociation of carbonates, XRD patterns showed predominant presence of calcium and magnesium oxides. In addition, pine ash which contains more

Treatment of Wastewater by Electrocoagulation Method…

31

manganese showed the presence of calcium manganese oxide and manganese oxide. Similarly, poplar, being richer in sodium, displayed weak peaks corresponding to sodium calcium silicate. It appears that when the ash is left standing in air, calcium oxide reacts with atmospheric water vapor to form calcium hydroxide. However calcium hydroxide is unstable at temperatures over 600oC. Table 6 also indicates that small amounts of potassium may be present as K2SO4 as the peaks corresponding to this compound become distinct at higher temperatures. Low temperature ash produced from the wood waste appears to contain predominantly calcium carbonate while at high temperatures the content changes to predominantly calcium oxide. What this Table shows is the close relationship of ash composition with combustion temperature. Many of these elements, when in solution, will behave as counter-ions. Wood ash leachate added to wastewater does the following: it hydrolyzes hydroxo-metallic positively charged ions are added to the wastewater medium the solution ionic strength is increased the solution electrical conductivity increases the positive hydroxo-metallic ions are adsorbed on the negative charge of the colloids surface, reducing the zeta potential to destabilization point the electrostatic distance between colloid particles is reduced and the energy barrier is overcome to allow agglomeration the presence of non-hydrolyzing counter-ions (Na+, Ca2+, Mg 2+) leads to the compression of the double layer (Figure 1) which leads to the reduction of the Zeta potential to van der Waals levels. with wood leachate, Al3+ and Fe2+ are also added and help neutralize the solution charge. They form precipitates that catch colloids in the flocs. these destabilized colloids and hydroxo-metallic complex by van der Waals forces lead to adsorption and flocculation.

Important Consideration of Wood Ash Leachate as Supporting Electrolyte One of the most important factors that need to be considered when using wood ash leachate as supporting electrolyte is the time required to allow leaching to take place and the ash to water ratio for leaching. In an experiment conducted on wood ash leaching, Etiégni and Campbell (1991) found that the total dissolved solids (TDS), K, Na, and Mg concentration increased linearly as the ash to water ratio increased (Figure 14). However, the percentage of ash dissolving did not change significantly, as approximately 10% of the ash dissolved at 50 g/L and 9% at 390 g/L.

32

Lazare Etiégni, K. Senelwa, B. K. Balozi et al. Table 5. Chemical composition of wood ash samples produced at different combustion temperatures given as the concentration in μg/g. (Source: Etiégni & Campbell, 1991). Element Aluminum Antimony Arsenic Barium Beryllium Bismuth Cadmium Calcium Cerium Chromium Cobalt Copper Iron Lanthanum Lead Lithium Magnesium Manganese Molybdenum Nickel Phosphorus Potassium Selenium Silicon Sodium Titanium Vanadium Ytterbium Zinc Zirconium Carbonate

Temperature (oC) 538 649 10,415 12,825 264 142 <20 <20 1,301 1,490 <5 <5 <30 <30 <2 <2 187,480 217,480 <40 <40 52 37 <3 <3 345 620 8,796 11,951 <10 <10 51 75 35 55 59,730 68,060 10,549 11,789 18 21 59 64 16,950 19,060 110,500 86,590 <50 <50 33,867 39,77 2,961 4,947 780 1,060 13 18 3 3 2,678 2,139 14 14 62.8 60.0

760 13,115 65 <20 1,640 <5 <30 <2 241,180 <40 49 <3 588 9,981 <10 200 70 76,970 13,489 26 151 21,720 47,720 <50 42,647 4,365 1,012 23 3 1,001 16 58.4

871 10,115 91 54 1,463 <5 <30 <2 216,080 <40 20 <3 380 8,219 <10 39 10 69,170 11,859 39 63 19,370 79,860 <50 31,997 1,643 777 17 2 177 15 54.9

982 13,415 94 74 1,601 <5 <30 <2 238,980 <40 39 <3 437 22,411 <10 43 30 80,470 16,609 56 497 21,030 51,970 <50 46,477 1,479 957 26 5 26 19 53.3

1093 11,015 208 59 1,530 <5 <30 <2 264,580 <40 622 <3 422 16,781 <10 685 10 89,100 21,680 192 8,979 25,460 4,026 <50 35,267 <100 796 234 3 55 23 51.4

As expected, the amount of TDS leaching from ash increased as pH decreased (Figure 15). For example, the percentage of TDS leached increased by 500% as the pH decreased from 13 to 5. Figure 15 show that, at any given pH, approximately 90% of the TDS were extracted within the first 30 minutes. Up to 58% of the total K and 65% of the total Na were leached with distilled water at pH=13 and a liquid to solid ratio of 20:1 (Figures 9 and 10). Very little magnesium was leached (0.003%). The change in electrical conductivity

Treatment of Wastewater by Electrocoagulation Method…

33

(EC) of the leachates followed the same trends as the TDS (Etiégni and Campbell, 1991). The reaction order of TDS leaching was found to be equal to n = 0.08 indicating that TDS was essentially independent of pH.

Table 6. XRD analysis of wood ash showing relative intensity of the strongest peaks (%). Compound

Pines 600oC

Aspen

Poplar

White oak

1300oC 600oC 1300oC 600oC 1300oC 600oC 1300oC 600oC 1300oC 600oC 1300oC

CaCO3

100

100

K2Ca(CO3)2

86

21

Ca(OH)2

24

34

40

*

3

MgO

8

100

26

4

49

4

19

100

16

100

100

2

*

CaO Ca4Mn3O10

100

100

100

100

11 8 100

100

21

Ca2MnO4

75

K2Ca2(SO4)3

12

Mg6MnO8

12

K2SO4

11

9

K2MgSi3O8

46

CaSiO3

*

Na2CaSiO4 Ca2SiO4

White Oak Bark Douglas Fir Bark

* 23

17 22

11

These findings have major implication on the use of wood ash leachate. They show that one can expect to extract most of the useful ion species within one hour. This should therefore constitute a critical design parameter, especially if the EC reactor must be a continuous one. One such reactor was recently patented and is shown in Figure 18 below. The most important fixture of this system is the ash tray and the mixing tank with three compartments. Baffles A1 and A2 prevent vortexes created by vigorous mixing in compartment 1. They are eliminated and allow laminar flow into compartment 2 and 3, thereby permitting ash particles to settle at the bottom of the mixing tank. The resulting mixture- ash leachate and wastewater- flows into the electrolytic tank where electrolysis takes place. Three variables are important here if the electrocoagulation system is continuous. The 1st one is the time required to allow the ash to leach which will be at least one (1) hour, the second will be the ash to water ratio and the third will be the detention time needed to permit maximum electrolysis before treatment. All these variables must be determined experimentally for each type of wastewater. If we assume plug flow: C = Co e-t/tR where: C = concentration of effluent Co = concentration of influent tR = Hydraulic detention time in the electrolytic tank.

(26)

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Lazare Etiégni, K. Senelwa, B. K. Balozi et al.

Figure 14. Leaching of solids at different ash concentration.

Figure 15. Leaching of TDS with time at different pHs.

Treatment of Wastewater by Electrocoagulation Method…

35

Sodium concentration (g/L)

6 5

Ash=40 g/l

4 Ash=30 g/l 3 Ash=20 g/l

2

Ash=10 g/l 1 Ash=5 g/l 0 0

5

10

15

20

25

30

Time (hr)

Potassium concentration (g/L)

Figure 16. Leaching of Na at different ash concentrations.

12 Ash=40 g/l

10

Ash=30 g/l

8 6

Ash=20 g/l

4 Ash=10 g/l 2 Ash=5 g/l

0 0

5

10

15

20

25

30

Time (hr) Figure 17. Leaching of K at different ash concentrations.

The hydraulic detention time would have been determined experimentally. The volumetric flow rate of the factory effluent will then enable one to size up the electrolytic tank: V (volume) = tR x Q (flow rate)

(27)

36

Lazare Etiégni, K. Senelwa, B. K. Balozi et al. Ash tank

+

-

Electrodes

Wastewater

A A 2 C2 1 C3

C1

Mixing Tank

Electrolytic tank

Leached ash

Leached ash draw off pipe Settled sludge

Sludge settling Tank

Sludge

Treated wastewater Source: Etiégni et al., 2004: International Patent: (PCT/ KE2005/000012). Figure 18. Process description of ELCAS for color removal from wastewater.

Because of the inherent requirement of leaching of wood ash, most reactors have been designed as batch. Other materials have also been investigated as potential sources of supporting electrolyte. They include: rock phosphate, ash from bagasse, coffee husk (Etiégni et al., 2007; Etiégni et al., 2009). Rock phosphate composition is shown in Table 7. The advantage with these materials is that they are produced on site, usually by the same factory that really needs them, at almost no cost. They have been successfully used to treat industrial wastewaters as shown in the picture below (Figure 19). The electrode configuration is depicted in Figure 20 below. This assembly with an electrode surface to volume ratio of 75 m2/m3 of effluent at a current density of 200 mA/m2 was successfully applied to completely decolorize wastewater from a coffee factory effluent. They can quickly neutralize acidic effluent and once they have been used, they can be safely applied on land as soil amendment. Rock phosphate produces less sludge than wood ash, but it is less effective for the reduction of power consumption (Etiégni et al., 2007). The only draw-back with wood ash leachate is that its tend to yield large amount of sludge (Orori, 2003; Barisa et al., 2009). The existence of these low cost supporting electrolytes is another tool at the disposal of

Treatment of Wastewater by Electrocoagulation Method…

37

wastewater treatment professionals. In terms of performance, the use of SE can only help the overall treatment of wastewater and more importantly reduce the overall power consumption.

Figure 19. Reactor used to treat coffee factory effluent.

The kinetic rate equation for representing the leaching of TDS or metallic ions from wood ash can be described by the following mth order reaction kinetics:

dC = −kCm dt

-----------------------------------(28) (26)

where C represents the TDS or the metallic ion concentration, m is the order of reaction, k is the reaction rate constant, and t is the time. For a first-order reaction, the above Eq. (28) becomes: ln (Ct/Co) = -k1t

(29)

The slope of the plot of ln Ct/C0 versus time gives the value of the rate constant k1, in t−1. Here, C0 is the initial concentration in milligrams per liter, Ct is the concentration value in g per liter at time t, and t is the time. The leaching reaction of wood ash can be described by a first order reaction where m = 1 and k = 0.007 in the case of potassium shown in Figure 21.

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Lazare Etiégni, K. Senelwa, B. K. Balozi et al.

Figure 20. Batch reactor with electrode configuration for color removal experiment.

Figure 21. Leaching reaction of K from ash.

Treatment of Wastewater by Electrocoagulation Method…

Figure 22. Leaching reaction of Na from ash.

Table 7. Chemical Analysis of Minjingu rock phosphate (Source: van Kauwenbergh, 1991). Compound P P 2O 5 CaO F Al2O3 CO32Na2O K2O MgO SiO2 Fe2O3 CO2 SrO BaO TiO2 S ClType of RP Reactivity NAC solubility

Concentration (%) Source (van Kauwenbergh, 1991) 10.9 38.3 6.9 1.2 2.6 12.5 1.3 Sedimenting Medium to high 23

NAC = Neutral ammonium citrate solubility

Concentration (%) 30.0 41.7 3.1 1.2 1.3 0.78 1.3 9.4 0.89 3.1 1.4 0.2 0.13 0.09 127 ppm -

In the case of Na (Figure 22), the reaction rate constant is much higher at 0.010

39

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Lazare Etiégni, K. Senelwa, B. K. Balozi et al.

ELECTRODES MASS LOSS Table 8 presents electrodes mass loss per year in a study of color removal from a pulp and paper mill effluent at several sampling points (SP1 to SP5) of an effluent treatment system (Orori et al., 2003). ELCON yielded the highest electrodes mass loss followed by ELCCA, then ELCAL and finally ELCAS at all sampling points as depicted in Table 8. Electrodes mass loss increased significantly (P 0.05) from SP1 to SP5 with all electrochemical color removal methods. Wood ash leachate seemed to reduce the overall mass loss of the electrodes, adding to another advantage of using this SE during electrocoagulation, because of the smaller time it took to attain total color removal, which saves or reduces electrode replacement cost. -1

Table 8. Mean electrodes mass loss year by electrochemical color removal methods. Sampling point SP1 SP2 SP3 SP4 SP5

-1

Electrodes mass Loss year ELCAS ELCCA Tones % Tones 1.63 0.024 5.52 2.87 0.041 9.67 3.68 0.053 12.7 4.66 0.067 15.39 5.98 0.086 20.38

ELCAL Tones 2.43 4.83 6.25 8.12 10.83

% 0.034 0.068 0.089 0.115 0.153

ELCON Tones 9.57 14.96 16.93 18.53 22.36

% 0.134 0.209 0.237 0.259 0.312

ELCAS

0.350

Percent electrodes mass loss per year (%)

% 0.079 0.138 0.181 0.220 0.291

ELCCA

0.300

ELCAL ELCON

0.250 0.200 0.150 0.100 0.050 0.000 SP1

SP2

SP3

SP4

SP5

Sampling points Figure 23. Electrodes mass loss comparison for all electrochemical color removal methods at all sampling points

Figure 23 shows the percent electrodes mass loss per year for all electrochemical color removal methods at these various sampling points. Percentage electrodes mass loss (<0.5%) was not statistically significant compared with the initial electrodes mass for all electrochemical

Treatment of Wastewater by Electrocoagulation Method…

41

color removal methods used. Electrodes mass loss increased significantly from SP1 to SP5 for all the EC methods as time for current flow increased. This was attributable to increasing original effluent color intensity from SP1 to SP5, which required longer time for color removal. Faraday‘s law can be used to relate the mass (m) of electrolytically generated aluminum going into solution to the operating current (I) and the run time (t). In this relationship, M is the atomic weight of aluminum or iron, z is the number of electrons transferred in the anodic dissolution (here z = 3 or z= 2), while F is Faraday‘s constant (96486 C mol-1). ItM m= zF

……………………………………………………… (27) (30)

Using this equation, the amount of coagulant or soluble metal delivered to the solution may be calculated.

SLUDGE PRODUCTION One of the biggest draw-backs of wood ash leachate as a supporting electrolyte during electro-coagualtion is its tendency to produce large amount of sludge. Orori et al. (2005) showed that ELCAS produced the highest sludge quantity compared to ELCAL and ELCCA, probably because of the contribution from wood ash leachate as can be seen in Tables 8 and 9. This has also been confirmed by other researchers (Barisa et al., 2009).

PRACTICAL APPLICATION OF WOOD ASH LEACHATE IN EFFLUENT TREATMENT There have been several applications of electrocoagulation experiments with wood ash leachate as supporting electrolyte. Orori et al. (2005) reported successful total color removal from a pulp and paper mill effluent using ELCAS process with close to 80% reduction of electric power consumption. This process has also been applied to remove color from Tea and coffee factory effluent.

OPERATING COST OF EC AIDED BY WOOD ASH LEACHATE AS SE The cost of the electrocoagulation process combined with wood ash or bagasse ash is determined by mainly two variables: the consumption of the sacrificial electrode and the electrical energy. These variables are the economic advantages of this method if one uses wood or bagasse ash which is almost free. But one needs to include the operating cost of sludge removal and compare the overall cost of ELCAS process with other methods. Operating cost + aCenergy + bCelectrode + CSE + C sludge removal

(31)

where CSE is the cost of supporting electrolyte which is almost negligible. It may include only some labor and transportation cost to the point of utilization. If we assume that the cost of

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Lazare Etiégni, K. Senelwa, B. K. Balozi et al.

electric power and that of electrode replacement will be cut by a least 50%, the operating cost of EC will be substantially reduced, if one was to use wood ash leachate. Table 9. Sludge Characteristics for ELCAL, ELCCA, and ELCAS at SP1. Sludge Characteristics Volume (ml/L) Dry weight (mg/L) Fixed solids (mg/L) Volatile solids (mg/L) Specific gravity Density

ELCCA 98.3600 5.6854 0.2471 5.4383 17.3005 0.0578

ELCAL 18.3600 0.2356 0 0.2356 77.929 0.0129

ELCAS 132.1400 6.5491 0.2854 6.2637 20.1768 0.0496

Table 10. Sludge Characteristics for ELCAL, ELCCA, and ELCAS at SP1 (Orori, 2003). Sludge Characteristics Volume (ml/L) Dry weight (mg/L) Fixed solids (mg/L) Volatile solids (mg/L) Specific gravity Density

ELCC A 98.360 0 5.6854 0.2471 5.4383 17.300 5 0.0578

ELCAL

ELCAS

18.3600 0.2356 0 0.2356 77.929

132.140 0 6.5491 0.2854 6.2637 20.1768

0.0129

0.0496

CONCLUSION Numerous research papers have shown that electrocoagulation method can be a more effective alternative treatment process for pollution abatement. However, certain factors such as electrode replacement and power cost are likely to make the method less affordable, although new sets of electrodes such as polypyrrole (PPy) and boron doped diamond (BDD) electrodes are becoming available on the market. BDD are expensive but more resistant and require higher current density. Fortunately, supporting electrolytes such as wood ash leachates, which are inexpensive and can make the process more affordable are available. A logical, systematic approach to a fundamental understanding of electrocoagulation with wood ash leachate as SE is clearly required. The design phase can then proceed on solid scientific and engineering knowledge. A large number of key mechanisms are dependent on a few operating parameters. The authors of this chapter have had quite a wealth of experience working with this SE. A trade-off between the competing factors such as sludge production, electrical conductivity and reduction of power consumption must be evaluated with respect to other optimum operating conditions.

Treatment of Wastewater by Electrocoagulation Method…

43

LIST OF ABBREVIATION EC SE ELCAS ELPHOS ELCON ELCAL ELCCA

= Electrocoagulation = Supporting electrolyte = Electrocoagulation combined with wood ash leachate = Electrocoagulation combined with rock phosphate leachate = Electrocoagulation alone = Electrocoagulation combined with alum = Electrocoagulation combined with calcium oxide/hydroxide

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Bhuptawat, H. K. & Chaudhari, S. (2003). Enhanced primary wastewater treatment with Moringa oleifera seeds. Proceeding of The International Conference on Water and Environment (WE-2003), Bhopal, India. Biswas, N. & Lazarescu, G. (1991) ―Removal of Oil from Emulsions Using Electrocoagulation‖, Int. J. Environ. Stud., 38, 65-72. Browning, E. (1996). “Chromium in: Toxicity of Industrial Metals”, second ed. Butter Worths, London, Calvo, L. S., Leclerc, J. P., Tnguy, G., Cames, M. C., Paternotte, G., Valentin, G., Rostan, A. & Lapicque, F. (2003) An Electrocoagulation unit for the purification of soluble oil wastes of high COD. Environmental Progress, 22(1), 57-65. Can, O.T., Kobya M., Demirbas, E. & Bayramoglu, M. (2006). Treatment of the Textile Wastewater by Combined Electrocoagulation. Chemosphere, 62, 181-187. Cenkin, V. E. & Belevstev, A. N. (1985). Electrochemical Treatment Of Industrial Wastewater. Eff. Water Treat. J., 25(7), 243-249. Chaudhary, A. J., Goswami, N. C. & Grimes, S. M. (2003). ―Electrolytic Removal Of Hexavalent Chromium From Aqueous Solutions‖. Journal of Chemical Technology and Biotechnology, 78, 877-883. Chiang, L. C., Chang, E. J. & Wen, C. T. (1995a). Indirect oxidation effect in electrochemical oxidation treatment of landfill leachate. Water Research, 29(2), 671. Chou, W. L., Wang, C. T. & Huang, K. Y. (2009). Effect of operating parameters on indium (III) ion removal by iron electrocoagulation and evaluation of specific energy consumption. Journal of Hazardous Materials In press doi:10.1016. Cossu, R., Polcaro, M. A., Lavagnolo, C. M., Mascia, M., Palmas, S. & Renoldi, F. (1998). ―Electrochemical treatment of landfill leachate: oxidation at Ti/PbO2 and Ti/SnO2 anodes‖ Environmental Science and Technology, 32(22), 3570. Do, J. S. & Chen, M. L. (1994). Decolorization of dye-containing solutions by electrocoagulation. Journal of Applied Electrochemistry, 24, 8:785-790. Eichhorn, E., Schneider, S., Speiser, B. & Würde, M. (1996). Recycling of the Supporting Electrolyte Tetra(n-butyl) ammonium Hexafluorophosphate from Used Electrolyte Solutions. Current Separations, 15, 2. Englehardt, J. D., Deng, Y., Polar, J., Legrenzi, Y. & Mognol, J. (2006). Options for Managing Municipal LandfillLeachate: Year 1 Development of Iron-Mediated Treatment Processes. Report #0432024-06. www.floridacenter.org Escobar, C., Soto-Salazar, C. & Toral, M. I. (2006). Optimization of the Electrocoagulation process for the removal of copper, lead and cadmium in natural water and simulated waste waters. J of Environ Management, 81, 381-391. Etiégni, L. & Campbell, A. G. (1991). Physical and chemical characteristics of wood ash. Bioresource Technology, 37, 173-178. Etiégni, L., Oricho, O. D., Ofosu-Asiedu, K., Senelwa, K. A, Surtan, K. G. & Omutange, S. E. (2007). Removal color from a craft pulp and paper mill effluent in Kenya using a combination of electrochemical method and phosphate rock. Journal of Water Science & Technology, Vol. 55, No. 6, 15-22. Etiégni, L., Oricho, O. D., Orori, O. B., Senelwa, K., Balozi, K. B. & Ofosu-Asiedu, K. (2009). Batch Reactor for Color Removal from a Coffee factory Effluent using a Combination of Electrocoagulation and different electrolytes. Journal of… In Press.

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Wang, P., Lau, W. C. I. & Fang, H. P. H. (2001a). Landfill leachate treatment by anaerobic process and electrochemical oxidation. Environmental Science, 22(5), 70. Wang, P., Lau, I. C.W., Fang, H. P. H. & Zhou, D. (2001b). Landfill leachate treatment with combined UASB and Fenton coagulation. Journal of Environmental Science and Health, Part A: Toxic/Hazardous Substances and Environmental Engineering, 35(10), 1981. WHO. (1984). Guidelines for Drinking -Water Quality. Vol. 2, Belgium, Macmillan. Xueming, C., Guohua, C. & Po, L. Y. (2002). Novel Electrode System for Electroflotation of Wastewater. Environ. Sci. Technol., 36(4), 778-783. Yildiz, Y. S., Koparal, A. S. & Keskinler, B. (2008). Effect of initial pH and supporting electrolyte on the treatment of water containing high concentration of humic substances by electrocoagulation. Chemical Engineering Journal, Vol. 138, No 1-3, 63-72.

In: Fluid Waste Disposal Editor: Kay W. Canton, pp. 49-81

ISBN: 978-1-60741-915-0 © 2010 Nova Science Publishers, Inc.

Chapter 2

APPLICATION OF SULPHATE-REDUCING BACTERIA IN BIOLOGICAL TREATMENT WASTEWATERS Dorota Wolicka* Institute of Geochemistry, Mineralogy and Petrology, Faculty of Geology, University of Warsaw, Żwirki i Wigury 93, 02-089 Warsaw, Poland,

1. INTRODUCTION 1.1. Trends in Environmental Biotechnology Human activity is strictly linked with the production of waste i.e. materials and substances that are undesired and cannot be used further. On the one hand these substances are natural to the environment, eliminated from further technological process by their uselessness (e.g. mining waste), or represent new products i.e. anthropogenic waste, being the by-product of industrial and agricultural activities. A separate group comprises municipal waste that is not linked with production but results from human dwelling. Utilization actions aiming at neutralizing and/or removal of waste are focused on substances that due to their existing or potential chemical activity may negatively influence the biosphere. Non-active substances represent alien elements in the natural environment, but due to their passive character, their utilization is concentrated on non-conflicting storage. Active pollutants influencing the natural environment penetrate it as gaseous emanations, fluids (sewage and effluents) and solids. Restriction of emission and removal of hazardous gaseous emanations should be conducted in places where they are formed. Imperfection of the applied technology or its lack results in atmospheric pollution. This problem can be of local (around industrial plants, e.g. chemical works, food processing plants, around farmsteads and stock farms), country or global range (emission of CO2, nitrogen compounds, gases hazardous to the ozone layer). Control of such pollutants beyond their source areas is difficult or even impossible.

*

Corresponding author: email: [email protected].

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Control of liquid pollutants such as sewage or effluents or solid waste is always focused on their chemical transformation in order to obtain end-products that are neutral to the environment. These methods are applied to sewage, whereas in the case of solid waste they may be used in the soil-water environment with essential and stable water supply, which as a solvent mobilizes solid compounds susceptible to leaching forming effluents within the dump sites or in their foreland, and facilitates the growth of microorganisms at the boundary between the solid and liquid phases. The main aim of each sewage treatment method is protection of the natural environment against unfavourable influence caused by introduction of such wastes. For many years attention was drawn on disturbance of the oxygen balance caused by presence of organic and ammonium compounds. Due to this fact, pollution treatment methods were dominated by methods ensuring distinct reduction of BOD5, COD, nitrification of ammonium compounds, and effective utilization of active sludge. Next, focus was drawn on eliminating inorganic compounds of phosphorus and nitrogen, degradation of non-biodegradable or poorly biodegradable compounds. The primary aim became, however, decreasing treatment costs, what is economically justified. These aims can be realized using many methods and on every stage of liquid waste utilization. New equipment or technology may be introduced, or those previously applied can be modified and optimized. It should be remembered that preventing environmental pollution does not begin at the stage of sewage treatment, but much earlier, and requires wide-range activities. The most correct attitude is preventing hazards at their source, particularly in the case of industrial waste. In the first place activities should be undertaken to diminish the volume and harmfulness of pollutants that are by-products of industrial processes or steps should be taken to work out a technology treating several types of sewage and/or waste in one process. The scale of this problem can be illustrated by the dairy industry, where the volume of sewage flowing out of a dairy plant reaches several cubic meters per 24 h and comprises 0.5 to 3 times the volume of processed milk. Similar quantities of sewage are produced by oil refineries, where 1 ton of processed oil results in about 10 to 18 cubic meters of refinerypetrochemical waste. Analogous results have been observed in the case of solid waste produced during technological processes, where the volume of waste distinctly exceeds the final product. For example, the industry of phosphorus fertilizers produces 5 t of waste (phosphogypsum) from 1 t of phosphorites. Such waste not only poses serious hazard to the biological equilibrium in the environment, but also distinctly increases the cost of the technological process, which should in this case also include utilization of liquid and solid waste. Treatment of different types of sewage requires application of many physical, chemical and mechanical methods. They include e.g. retaining of suspensions from sewage on screens, sieves and in settling tanks, neutralization of acidic or alkaline sewage, adsorption of sewage components on relevant adsorbents, coagulation of non-subsiding suspensions, or extraction of sewage components by relevant solvents. Although commonly applied, these methods are not environment friendly, do not entirely solve the problem of waste neutralization, and often change only the physical-chemical composition or form of the waste. In such cases, anthropogenic waste is produced, which after many years of active production in the industrial plant may even become anthropogenic deposits. They are often the source of many rare elements e.g. elements from the lanthanoid or actinoid series (rare earths), which occur in considerable quantities in the waste, but cannot be recovered due to lack of efficient process.

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Thus it seems crucial to search for pro-ecological methods focused on the utilization of industrial sewage and waste that are hazardous to the natural environment. To such methods belong biological methods that are used in a wide range in the case of some types of organic sewage. All these methods use microorganisms to remove organic and some inorganic compounds from sewage. They decrease the volume of pollutants in the sewage and retain correct parameters determined by norms for sewage water introduced into surface waters. Selection of the appropriate method of sewage treatment depends on the type of sewage, its composition, volume, as well as degree of pollution of the water reservoir, to which the refined sewage will be introduced. These methods can be variously modified, and in some cases multistage sewage treatment is carried out with application of different methods. Biological methods may be applied only to the treatment of sewage, in which the concentration of toxic compounds does not hamper the incubation of microorganisms. Due to this fact, continuous cultures are applied in biological treatment plants, which allow obtaining the maximal incubation speed of microorganisms through continuous supply of fresh medium and removal of metabolism products. In a classical continuous culture (chemostat), the time of sewage flow should correspond to the microorganism growth speed, which is determined experimentally by change of flow speed from the moment when the biomass stable level is attained. Recently, search is focused on methods that would allow simultaneous biodegradation of several wastes. Concurrent biodegradation of two industrial wastes seems an interesting issue from the economy of the process. Costs linked with simultaneous biodegradation of two hazardous industrial wastes are always lower than for each of them separately. Biotransformation of phosphogypsum in the environment of organic sewage as a liquid state to its dissolution may be an interesting example. This trend in environmental biotechnology results in a large number of reports focused on treatment of high-sulphur sewage using sulphidogenesis methods (Lens et al., 1998) or biotransformation of solid waste in organic sewage environments. There are also single publications on the application of sulphidogenesis on the treatment of sewage after enrichment with phosphogypsum formed as a by-product in many industrial branches (Deswaef et al., 1996; Kaufman et al., 1996; Wolicka et al., 2005; Wolicka & Kowalski, 2005; Wolicka & Kowalski 2006a; Wolicka & Kowalski, 2006b; Wolicka, 2008b; Wolicka & Borkowski, 2008; Wolicka & Borkowski, 2009). Concluding, the most aggressive in the natural environment are liquid substances and they pose the most serious hazard to the natural environment.

1.2. Advantages and Disadvantages of Anaerobic Processes Biological treatment of various anthropogenic waste commonly applies aerobic methods, mainly due to the fast rate of the processes. However, the composition of various wastes often implies the application of anaerobic processes in the treatment procedure. In comparison to aerobic methods, this application has many advantages. First of all, anaerobic methods do not require expensive aeration, what often constitutes high costs for the whole processing plant. Further, it is estimated that only 6% COD are transferred into excessive sludge, which in many processing plants using activated sludge methodology generates additional waste. Very commonly the chemical composition of the excessive sludge does not allow it to be used e.g. in agriculture, but it can be utilized in an additional process in anaerobic conditions.

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Nowadays, sulphate reducing bacteria (SRB) are becoming more frequently applied in the biodegradation of anthropogenic waste. These bacteria in course of anaerobic respiration produce hydrogen sulphide, which can bind heavy metals in poorly soluble and non-toxic sulphides of metals. This is one of the many advantages of sulphidogenesis application in environmental biotechnology. Additionally, due to the toxic activity of hydrogen sulphide, SRB may eliminate various microorganisms from the environment, including pathogenic forms, what causes their domination in a sulphate rich environment. The bacteria may be utilized during the biodegradation of two industrial wastes, of which one may be solid waste as the sulphate source, and the second – liquid waste as the carbon source. Application of such process allows simultaneous biodegradation of two arduous industrial wastes, thus shortening the biodegradation time. Moreover, post-culture deposits generated in this process i.e. carbonates and/or calcium phosphates can potentially be utilized in agriculture. An ideal example is the biotransformation of phosphogypsum. It should be remembered, however, that anaerobic methods are not devoid of disadvantages, including: 1. difficulties in retaining the concentration of a particular microorganism group in the bioreactor due to the easy formation of symbiotic relationships, e.g. SRB easily form microbiological consortia with metanogenic archea; 2. anaerobic processes are very sensitive to pH and temperature changes and oscillation of hydraulic and substrate loading; 3. due to the multistage biodegradation process in anaerobic conditions, the duration of complete mineralization of organic compounds is much longer than in the case of aerobic processes; 4. the distinct disadvantage of common SRB application is toxic hydrogen sulphide produced during metabolic processes. This fact obliges to introduce an additional stage, during which hydrogen sulphide is oxidized e.g. to elemental sulphur. Additionally, bacteria of the genus Desulfovibrio influence the biocorrosion process, which is very often the case of destroyed metal elements of the hydraulic installation. This is linked with the presence of sulphates and SRB activity, what results in the reduction of oxygenated sulphur compounds, and the formation of sulphides according to the following reaction: 4Fe + SO42- + 2H2O + 2H+ → FeS↓ + 3Fe(OH)2 Other metals that are present in the alloy e.g. Cu, Zn, Ni or Cr may also take part in the reaction.

2. ECOPHYSIOLOGY OF SULPHATE REDUCING BACTERIA (SRB) 2.1. Environments of Occurrence Sulphate reducing bacteria (SRB) are heterotrophs and absolute anaerobes. They utilize sulphates, as well as other partly oxidized sulphur compounds (sulphites, tiosulphites and

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tetrationates), and elemental sulphur as the final electron acceptor in the respiration processes (Postgate, 1984; Gibson, 1990). Electron donors for this microorganism group are organic compounds such as e.g. alcohols, carboxylates, phenols, aliphatic and aromatic hydrocarbons, amino acids and some carbohydrates. Diverse SRB physiology influences their distribution in the natural environment as well as in anthropogenic environs e.g. polluted by crude oil and oil products (Wolicka& Borkowski, 2007a; Wolicka, 2008a). The presence of SRB has been noted in aquatic and terrestrial environments (Hao et al., 1996). They occur in soils, deposits of fresh water and marine reservoirs, in silts at the moutha of river deltas etc. (Chi Ming So & Young 1999), thermal springs and in geothermal regions, in crude oil, refining and petrochemical waste, natural gas intakes and on corroding steel (Hao et al., 1996). They may be found in all types of bioreactors purifying sewage in anaerobic conditions, from which they can be isolated (Przytocka-Jusiak et al., 1997; Baena et al., 1998, 1999, 2000; Hernandez et al., 2000). The most characteristic environments of SRB occurrence are marine deposits (Bak & Widdel, 1986; Szewzyk & Pfennig, 1987; Lovley et al., 1995; Aeckersberg et al., 1998; Caldwell et al., 1999; Kniemeyer et al., 2003), in which sulphate concentration reaches averagely 28mM (Wit, 1992), as well as oil fields and crude oil reservoirs (Voordouw et al., 1996; Mueller & Nielsen, 1996; Jenneman & Gevertz, 1999; Magot et al., 2000; Wolicka, 2008a). Their presence has also been noted in environments polluted by crude oil and oil products, dairy work sewage, whey, refining-petrochemical waste and distillery decoctions (Wolicka & Kowalski, 2005; Wolicka, 2006; Wolicka, 2008b; Wolicka & Borkowski, 2009). Additionally, SRB always accompany crude oil and were for a long time considered as microorganisms indicative of oil deposits (Postgate, 1984). Suggestions pointing to the presence of SRB in brines from oil fields come from 1926. They are the first attempt to explain the permanent presence of sulphides in crude oil reservoirs (Jenneman & Gevertz, 1999). SRB are the most commonly isolated group from oil fields and their groundwater (Rueter et al., 1994; Mueller & Nielsen, 1996; Aeckersberg et al., 1998; Wilknes et al., 2000; Magot et al., 2000; Rozanowa et al., 2001). Although SRB are considered as absolute anaerobics, their presence has also been noted at the boundary between the oxygenated and anaerobic zone in sediments, or even within the oxygenated zones. The presence of SRB is often recognized in the environment due to the presence of the characteristic odour of hydrogen sulphide as well as due to black colouring being the effect of precipitation of poorly soluble metal sulphides (Postgate, 1984; Gibson, 1990).

2.2. Metabolic Processes Carried by SRB Sulphate reducing bacteria (SRB) belong to absolute anaerobes, utilizing oxide compounds of sulphur as the final electron acceptors. They gain energy from oxidation of easily accessible organic compounds, and the electrons detached from the substrate are transferred on the sulphate according to the formula: organic compounds + SO42- → H2O + CO2 + HSThe preferred carbon sources for SRB are low-molecule organic compounds such as organic acids e.g. lactic, formic, pyruvic, and malic acids; volatile organic acids e.g. acetic acid,

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and alcohols e.g. ethanol, propanol, methanol, and butanol. Some SRB species are known to utilize amino acids as the sole carbon source: Desulfovibrio aminophilus (Baena et al., 1998), Desulfobacterium vacuolatum (Rees et al., 1997), and Desulfovibrio mexicanus (HernandezEugenio et al., 2000). Some species, e.g. Desulfotomaculum antarcticus may utilize glucose as the sole carbon source, but this is a rare case in SRB (Fauque et al., 1991). Rather common in turn is the utilization of aromatic and aliphatic compounds (Widdel & Bak, 1992). All organic compounds that represent the optimal carbon source for SRB are the products of fermentation formed during anaerobic biodegradation of carbohydrates, proteins and lipids (Fauque et al., 1991; Hao et al., 1996). This results from the fact that SRB do not produce hydrolytic enzymes and become involved in the anaerobic biodegradation of organic matter in the last stage. The only exception is the archeon Archeoglobus fulgidus, which can produce hydrolytic enzymes. Many papers devoted to the anaerobic biodegradation of crude oil contain information of the utilization of SRB in this process, as well as in the biodegradation of oil products both in refinery-petrochemical waste and in bioremediation of soils polluted by crude oil and oil products. Biodegradation of n-hexane, n-octane and n-decane by various SRB has been described by Gieg & Suflita (2002). Aeckersberg et al. (1998) described two mesophilous strains Hxd3 and Pnd3 utilizing n-alkanes C12-C20 and C14-C17, whereas So & Young (1999) described a mesophilous strain AK-01 utilizing n-alkanes C13-C18. Tribe TD3 was able to grow on media with n-decane and n-alkanes C6-C16 (Reuter et al., 1994). Utilization of aromatic compounds such as benzene, toluene and xylene in SRB cultures has been noted by Edwards & Garbic-Galic (1992), Beller et al. (1992), Edwards et al. (1992) and Ball & Reinhard (1996). Benzene, toluene, ethylbenzene and xylene (orto-, meta-, paraxylene) were utilized by thermophilous sulphidogenic consortia ALK-1 and LLNL-1 described by Chen & Taylor (1997). Benzene decomposition by thermophilous SRB was noted by Lovley et al. (1995), Przytocka-Jusiak et al. (1997) and Caldwell et al. (1999). Desulfobacula toluolica - Tol2 (Rabus et al., 1993) and Desulfobacterium cetonicum (Harms et al., 1999) are able to utilize toluene as the sole carbon source. The PRTOL 1 tribe isolated from soil polluted by petrol used toluene, phenyl propionate, phenyl acetate, benzoaldehyde, benzoate, p-cresol and p-hydroxybenzoate as the sole carbon source (Rabus et al., 1993). Tribe EbS7 completely oxidizing ethylbenzene was described by Kniemeyer et al. (2003). Tribes oXyS1 and mXyS1 are capable to utilize o-xylene (2%) and m-xylene (2%) as the carbon source (Harms et al., 1999). Both tribes use also toluene (2%) and benzoate. Many SRB species are known to utilize also other organic compounds such as phenol, catechol, cresol or indole. For example Desulfobacterium phenolicum (Bak & Widdel, 1986) is able to biodegrade phenol as the only carbon source, Desulfobacterium catecholicum – catechol (Szewzyk & Pfennig, 1987), Desulfobacterium indolicum (tribe In04) biodegrades indole (Bak & Widdel, 1986), and Desulfobacterium cetonicum biodegrades m-cresol (Galushko & Rozanova, 1991). The mesophilous SRB species Desulfotomaculum gibsoniae used phenol, catechol, metylocatechol, m-cresole and p-cresole as the sole carbon source and sulphates, sulphides and tiosulphates as electron acceptors (Kuever et al., 1999). The sulphidogenic consortium comprising Desulfovibrio desulfuricans A, D. desulfuricans B., D. gigans and D. vibrio was able to biodegrade 1,3,5-trinitrobenzene, hexahydro-1,3,5-trinitro-1,3,5-triazine and -1,3,5,7tetranitro-tetrazocine (Boopathy et al., 1998), whereas tribe BRS- HHQ20 utilized 1,2,4trihydroxybenzene (Reichenbecher & Schink, 1997).

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Table 1. Metabolic reaction of SRB and the content of released energy (after Hao et al., 1996). reaction 3 lactate → 2 propionate + acetate + HCO3- + H+ 2 lactate + SO42- + H+ → 2 acetate + 2CO2 + HS- + 2H2O 4 propionate + 3SO42- → 4 acetate + 4HCO3- + 3HS- + H+ acetate + SO42- → 2HCO3- + HSacetate + 4S + 3H2O → 4H+ + HCO3- + 4HS- + CO2 4H2 + SO42- + CO2 → 3H2O + HS- + HCO3H2S + SO42- + H+ → HS-+ 4H2O

∆G0’ (kJ/mol) -165 -189 -151 -60 -24 -152 -172

Over 40 genera of microorganisms taking place in the dissimilation reduction of sulphates have been subdivided into two main groups depending on the product of decomposition of the organic substrate (Brock & Madigan, 2006): I group – bacteria that have the ability to utilize fatty acids with an even carbon number: lactate, pyruvic amid, ethanol, and some fatty acids as electron donors, but they oxidize them to acetate reducing sulphate (VI) to hydrogen sulphide, thus carbon dioxide is not produced. This type of metabolism is characteristic of e.g. Desulfomonas, Desulfovibrio, Desulfobulbu and Desulfomicrobium. II group – bacteria that completely oxidize such compounds as e.g. fumaric acid, acetate, lactate and oxaloacetic reducing sulphates (VI) to sulphides. As a result CO2 and H2O are produced. To this group belong e.g.: Desulfosarcina, Desulfonema, Desulfococcus, Desulfobacterium and Desulfotomaculum. SRB acquire energy in two main processes: phosphorylation linked with transport of electrons – accompanying sulphate reduction, as well as during oxidation of organic substrates and the accompanying substrate phosphorylation (Hao et al., 1996). The assumption that phosphorylation takes place at the level of electron transport is confirmed by the presence of cytochromes and iron-sulphur proteins and by the distinct energy profit of the process. In SRB that have the ability to use lactate and/or pyruvic acid, ATP originates in oxidation phosphorylation as well as during substrate phosphorylation, i.e. oxidation of lactate through pyruvic acid to acetate, carbon dioxide and hydrogen (Brock & Madigan, 2006). Although SRB are able to utilize a wide variety of organic compounds, their reactions with different electron donors lead to the formation of HCO3- and HS- ions (Table 1) (Hao et al., 1996).

Factors influencing SRB growth Besides easily accessible carbon sources and the presence of oxidized sulphur compounds, many factors influence the life and growth of SRB. These physical and chemical factors include: concentration of dissolved oxygen, temperature, pH, Eh, and presence of accompanying microflora. Studies by Hao et al. (1996) have shown that concentration above 1.0 mg O2/l leading to the increase of the redox potential, in effect inhibits SRB activity. On the other hand, some tribes such as e.g. Desulfovibrio desulfuricans, D. vulgaris, D.

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desulfodismutans, Desulfobacterium autotrophicum, Desulfolobus propionicus and Desulfococcus multivorans may survive at oxygen concentrations below 0.5 mg/l, even utilizing it as the electron acceptor (Dilling & Cypionka, 1990). Another important factor influencing the activity of SRB is temperature. Most tribes and sulphidogenic assemblages display the optimal activity within the range 28–32 C. There are also species that survive in higher or lower temperatures; e.g. some tribes of Desulfobacter prefer temperature in the range 24–28 C, and Thermodesulfobacterium commune – about 70 C (Hao et al., 1996). The optimal growth temperature of archaea such as Archeoglobus fulgidus and A. profundus isolated from hydrothermal waters is at 80 C. It should pointed out, however, that temperature exceeding 45 C may be killing for SRB. Similarly as temperature, pH of the environment may represent a factor inhibiting SRB growth. It is commonly considered that these organisms prefer environments with a neutral reaction, whereas pH below 5.5 and above 9.0 inhibits their growth (Hao et al., 1996). On the other hand, a number of papers point to the occurrence of SRB in acidic mine waters, where pH is about 3, or in exploitation workings where this value fallows down to 2 (Wolicka & Borkowski, 2007b). Strong inhibiting and even toxic activity of SRB has been observed in the case of H2S as well as HS- and S2- ions. They react with cytochromes and iron of ferredoxines, as well as with other intracellular compounds containing this element, retarding the system of electron transport (Hao et al., 1996). Other known inhibitors of sulphate reduction are selenite ions (SeO42-), antagonists of sulphate reduction, as well as molibdate ions (MoO42-), which are structural equivalents of SO42- ions and cause the cellular ATP content. Recently, excessive iron ions in the environment are also considered to inhibit sulphate reduction. SRB similarly as most microorganisms are sensitive to bacteriocides and bacteriostatics (Hao et al., 1996). Some bacteria belonging to SRB are able to utilize electron acceptors other than SO42- i.e. 2SO3 and S2O32-. The utilization of tiosulphates and sulphites as the final acceptors allows bacteria to obtain energy indispensable for growth, because they do not reduce sulphates. A representative of this group is Desulfovibrio sulfodismutans, which has the ability to reduce each of these compounds according to the following reactions (Hao et al., 1996): S2O32- + H2O → SO42- + HS- + H+ G0‘ = -21.9 kJ 4SO32- + H+ → 3SO42- + HS- G0‘ = -235.6 kJ In anaerobic conditions elemental sulphur may also represent the final electron acceptor. This process is known as sulphur respiration and is not as common as sulphate reduction. A representative of this group is Desulfuromonas acetoxidans that utilizes acetate (rarely ethanol, propanol) as the carbon source and electron donor, and conducts its complete oxidation to CO2 according to the following reaction (Bothe & Trebst, 1981). CH3COOH + 2H2O + 4S0 → 2CO2 + 4H2S G0‘= -5.7 kcal/mol Electrons detached during oxidation of organic compounds are transported in the respiration cycle on cytochrome c7, next on protein 4Fe-S, and then reach the final electron

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acceptor – elemental sulphur. The ability to reduce elemental sulphur has been noted also in other microorganisms, among which occur: Sulfospirillum – utilizing most frequently H2 as the electron donor. Desulfurella – thermophilous bacteria utilizing acetate as the electron donor. Campylobacter – not able to reduce sulphates but reducing sulphur, sulphites, tiosulphates, nitrates as well as fumaric acid, utilizing acetate as the electron donor. Pyrodictium – a thermophilous archeon able to utilize diatomic hydrogen as the electron donor, and elemental sulphur as electron acceptor. Particular SRB species differ in several features, including: cell shape, mobility, occurrence, preferred electron donors, complete or incomplete oxidation of organic compounds, content of GC pairs in DNA, formation of spores, presence of desulphoviridine, cytochromes, and optimal growth temperature (Gibson, 1990). Based on rRNA analysis, SRB have been subdivided into four groups: 1. 2. 3. 4.

gram -, mesophilous SRB; gram +, SRB generating spores; thermophilous SRB; thermophilous archaeabacteria reducing sulphates.

2.3. SRB Isolation and Culture Methods Till 1984, SRB were considered to be dominated in the natural environment by other microorganisms. Therefore, SRB were commonly isolated on media containing an easily accessible carbon source, e.g. lactate, pyruvic acid or ethanol, and sodium sulphate, well soluble in water, as the electron acceptor (Postgate, 1984). However, the last twenty years have brought many reports on the fact that SRB are microorganisms that have the ability to biodegrade a wide spectrum of organic compounds. In due course, SRB began to be isolated on media containing compounds that are the main pollutants in the environment. Thus, active SRB communities are isolated from environments polluted by substances that are subject to treatment. For example, anaerobic SRB communities can be isolated from refinerypetrochemical sewage, soils polluted by oil products, or other organic sewage, and are then used in the anaerobic utilization process of these wastes. Very soluble sulphates such as sodium sulphate were considered the optimal electron acceptors in the sulphidogenic process. There are also reports indicating the fact that poorly soluble sulphates such as bassanite (CaSO4 x 0,5H2O), gypsum (CaSO4 x 2H2O), anglesite (PbSO4), or barite (BaSO4) may also become an easily accessible electron acceptor for SRB (Karnachuk et al., 2002). During multiplication of anaerobic microorganism communities, typically two methods of SRB selection are applied: the ―microcosms‖ method and multiplication on agar medium (Figure 1). Reproduction of selected microorganism communities should take place in strictly anaerobic conditions. Therefore, liquid media are often supplemented with compounds decreasing the redox potential, e.g. cysteine, sodium tioglicolate, or sodium sulphide, as were as indicators of oxygen level such as resaurine.

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Figure 1. Methods of isolation of sulphate reducing bacteria.

The ―microcosms‖ method is considered one of the best multiplication methods of anaerobic bacteria communities from the environment. This fact was observed by Lesage et al. (2000) and Wolicka (2006, 2008a). During SRB multiplication and isolation, enrichment is based on the addition of solutions of sulphates and particular organic compounds that comprise the main pollutants, e.g. casein in the case of dairy sewage and phenol in the case of refinery-petrochemical waste. The method was applied by Warren et al. (1999) during isolation of acetotrophic methanogenes; Lesage et al. (2000) during multiplication of anaerobic bacteria communities able to utilize polycyclic aromatic hydroharbons; Zhang Weixian & Bouwer (1997) during isolation of microorganisms biodegrading benzene, toluene and naphthalene from sludge; Kleikemper et al. (2002) and Wolicka & Kowalski (2006b) from slops, renifery-petrochemical waste (Wolicka et al., 2005) and dairy sewage (Wolicka, 2008b). Multiplication of anaerobic bacteria is conducted in tall and narrow vessels, what restricts oxygen diffusion. Air penetration to the culture is shut off by e.g. a rubber cork. The cultures should not be opened during incubation, and samples for determinations should be collected by a syringe needle inserted in the cork. In monospecific cultures of obligatory anaerobic bacteria additional precautions are taken, such as addition of compounds decreasing the redox potential to the medium (cysteine hydrochloride or sodium tioglicolate) or purging the culture by neutral gases. In polyspecific cultures from natural environments, anaerobic conditions are gradually developed during usage of oxygen by aerobic bacteria. Many difficulties may be dealt with during isolation of pure SRB strains. SRB are absolute anaerobes; this fact requires anaerobic conditions during the entire isolation process,

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what is very difficult to achieve and obliges the application of e.g. neutral gases to decrease the redox potential of the culture. A different issue is the multiple passaging of the culture, which is indispensable in the process but at the same time causes oxygenation of the medium. SRB isolation from agar media is facilitated by the black colouring of the cultures on media containing iron caused by formation of iron sulphides. Another difficulty comes from the fact that SRB grow in strict physiological and spatial relationships with other groups of bacteria, e.g. they form consortia with methanogenic archaea. Decomposition products formed by one group within the consortium become the substrate for other bacteria groups taking part in the anaerobic decomposition of organic matter. During SRB identification it often turned out that strains considered pure, in fact comprised a microorganism community that was not possible to sub-divide, or after isolation each of the components was not capable of independent growth.

III. APPLICATION OF SULPHATE REDUCING BACTERIA IN BIOLOGICAL TREATMENT OF WASTEWATERS 3.1. Characteristics of Wastewater in Anaerobic Treatment Prior to application of biological treatment methods on sewage, various physical and chemical analyses are made in order to characterize the sewage and select the appropriate treatment method. The analyses determine: pH, temperature, colour, smell, turbidity, suspension content, COD, BOD, and content of inorganic compounds. Important data include also the volume of sewage that should be treated in a given time period, as well as the sewage discharge method, i.e. whether the sewage will flow steadily or periodically to the treatment apparatus. Sewage is produced in many branches of industry. According to the definition, sewage represents water that is polluted beyond its natural reservoirs, thus it can be assumed that precipitation water (rain, hail, snow) polluted on streets in towns and cities, as well as on the surface of fields also becomes sewage. Water drawn from surface reservoirs for consumption, sanitation or production is polluted in households and various types of service and production plants. The content and quality of sewage depends thus on the industrial development and intensity of economical activities in a given country or region, as well as on climate. Sewage that is planned to be treated during sulphidogenesis should fulfil several prerequisites. It should contain oxygenated sulphur compounds such as e.g. sulphates that are the main electron acceptor for the SRB, as well as easily accessible organic compounds. Lactate has been previously considered as the optimal carbon source for SRB, however application of this compound as a carbon source in high-sulphur sewage is economically not feasible. Therefore the search began for other carbon sources available to SRB and constituting the main pollutants in organic sewage. Organic sewage contains organic compounds that are either easily biodegradable by microorganisms, or, on the contrary, are poorly or very poorly biodegradable. In general, during sulphate reduction organic acids and alcohols are easily biodegradable, whereas e.g. aliphatic and aromatic hydrocarbons, pesticides, herbicides and xenobiotics are poorly biodegradable. Sewage polluted by organic compounds susceptible to biodegradation is

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produced mainly by food industry, e.g. dairy plants, sugar plants, fruit and vegetable processing plants, meat processing plants, breeding plants, and slaughter houses. Dairy sewage contains high volume of organic compounds and BOD. The main constituents of this sewage are saccharides, fats and proteins that compose milk. This sewage undergoes biodegradation easily, after which the environment becomes acidified and hydrogen sulphide, toxic for most microorganisms, is produced. The sewage cannot be introduced into the receiver without treatment. The volume of sewage produced by a single dairy plant reaches several thousands of cubic meters per 24 h. The composition of dairy sewage resembles highly diluted whole milk, with dissolved lactose and protein (casein) in an identical proportion as in milk. Based on the analysis of milk composition used in the processing plant and the production profile, the sewage composition can be assumed. The volume of sewage formed in dairy plants comprises from 0.5 to 3 times the volume of processed milk. Typically, 1 g BOD5 in sewage is considered to correspond to 9 g BOD5 in milk. The BOD5 of sewage reaches 450–5800 mg O2/l (averagely about 1800 mg O2/l). The total volume of dairy sewage comprises strongly polluted production sewage formed during flushing of products or washing of vessels, equipment, facilities, as well as slightly polluted cooling waters (comprising 60–90% of sewage volume). Sewage may also contain low quantities of flavouring matter, gelling and clarification agents, etc., as well as disinfectants, cleaning and degreasing agents, and agents removing limescale from utensils. Sewage reaction varies between 7.0 and 8.8, with the exception of acidic sewage from production of casein and selected cheeses, which varies between 5.5 and 6.5. Due to the content of lactose, dairy sewage is easily biodegraded, and its pH falls down to very low values. The composition of sewage from particular dairy plants may vary considerably depending on the type of manufactured products (Tab. 2, 3, 4). The table 2 presents the typical composition of dairy sewage from a plant producing cream, cottage cheese and kefir. The ballast of dairy sewage is more variable when it encompasses buttermilk, whey and skimmed milk. The application of some sewage constituents e.g. whey to feed breeding animals, or buttermilk to produce beverages, may distinctly influence this ballast. Whey contains about 72% lactose, 10% protein, 0.5% fat, as well as mineral compounds and vitamins. In the dairy industry it is formed as a by-product during cheese or casein manufacture. One volume of produced cheese results in almost 10 volumes of whey, which is a crucial problem in cheese manufacture (Kutera & Talik, 1996). Table 2. The typical composition of sewage from plant producing milk products. Determinations Total dry mass Mineral dry mass Organic dry mass BOD5 COD Total nitrogen Phosphorus

Content [g/m3] 1273 626 647 587 1048 33 13

Determinations Potassium Calcium Magnesium Sodium Sulphates Chlorides pH

Content [g/m3] 41 77 15 93 258 95 7.3

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Casein and lactose are not optimal carbon sources for SRB. Instead, products of their hydrolysis may become electron donors for SRB, as these bacteria do not produce hydrolytic enzymes and therefore they take part in the process of organic matter biodegradation at the level of volatile fatty acids (Mizuno et al., 1998). Due to this fact, utilization of proteins and disaccharides is a rare case in SRB. Table 3. Typical composition of sewage formed during washing of pipelines for milk transport. Determinations pH Oxidization [mg O2 / dm3] COD [mg O2 / dm3] BOD5 [mg O2 / dm3] ether ekstract [mg / dm3] Suspensions [mg / dm3]

Content 9.6 290 1200 515 160 480

Table 4. Typical composition of UHT milk. Determinations Lactose Fats Proteins Mineral salts Water

Share 4.7% ~3.2% ~3% 0.8% to 100%

Sewage polluted by organic compounds that are not easily biodegraded includes sewage produced in e.g. oil refinery and petrochemical plants, pulp and paper industry, textile industry, leather industry, as well as plant and animal utilization plants. Refinery-petrochemical sewage contains from about 0.3 to 2% of organic compounds from crude oil, such as hydrocarbons, alcohols, aldehydes, esters, alkali, acids and their salts from the refining-petrochemical industry, as well as oily pollutants (with a various degree of emulsification) and tars. In petrochemical plants sewage is formed in the process of raw material cleaning, as well as during the manufacture of many chemical products, half-products and usable products from crude oil and its fractions and from earth gas (most commonly distillation and rectification processes). Sewage composition depends on the type of production in a given petrochemical plant. The main pollutants in the sewage are hydrocarbons, alcohols, aldehydes, phenols, esters, alkali, acids and theirs salts. Other common constituents of petrochemical sewage include phenol, ethylene, propylene, butadiene, acetone, glycol, plastics, synthetic rubber, epoxy resins, and surface active agents. Petrochemical sewage is commonly treated jointly with refinery sewage. Petroleum processing includes thermal processing comprising distillation, rectification and cracking. In modern refinery plants processes of further fuel refinement are carried out, including petrol reforming and hydrorafinery (desulphurization) of diesel oil. The first phase of crude oil processing comprises removal of gas in gas separators, mechanical pollutants by

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water flushing and sedimentation in settler tanks, whereas emulsions are broken in electromagnetic field, by heating or addition of deemulsifiers. After these initial stages, crude oil is distilled by separation of hydrocarbons into fractions characterized by similar boiling temperatures. At normal atmospheric pressure are produced: petrol, heavy petrol, naphtha, diesel oil and oil fuel. Distillation of fuel oil at lower pressure leads to the production of diesel oil, light, medium and heavy petrol (paraffin oils). Hydrocarbon fractions obtained during distillation are refined in order to obtain purified products. Modern oil refinery methods are based on the adsorption of pollutants or application of selective dissolvents that dissolve the pollutants not dissolving the oil. The most commonly applied dissolvents are phenol and furfuryl alcohol. The content of sewage produced in oil refineries depends on the quality of oil and degree of its processing, and varies from 10 to 18 cubic meters per tonne of processed crude oil. Sewage is produced during: -

washing of facilities (high volumes of oil and oil products); during flushing of oil and dewatering of raw oil (fatty and naphthene acids, phenols, inorganic salts, mainly NaCl); during crude oil distillation and cracking (hydrogen sulphide, thiols); during refinery: acidic sewage (sulphuric acid, resins) and alkaline sewage (sodium hydroxide, sodium hydrosulphide, sodium salts of fatty acids, phenolates and emulsions of oil products)

The colour of sewage varies from milk white to dark brown and it has a specific odour (Table 5). The main pollutants in sewage from refining plants include naphtha, sulphides, thiols, phenols, fatty and naphthene acids, naphthene sulphoacids, mineral oils, aliphatic and aromatic hydrocarbons (often chlorinated), aldehydes and alcohols formed during oxidation of hydrocarbons, and suspensions. Naphtha occurs in sewage in forms of emulsions or surface films. Table 5. Characteristics of sewage formed during naphtha refinery. Determinations Temperature Odour Transparency Reaction Suspensions Oxidization Phenols Hydrogen sulphide Oils

units o C descriptive cm pH mg/l mg O2/l mg/l mg H2S/l mg/l

results 15–63 naphtha and H2S 1–4 6.5–8.4 200–4200 650–1200 50–150 10–30 20–80

Inorganic sewage that may be biodegraded during sulphidogenesis results from e.g. production of artificial fertilizers, mines, manufacture of various inorganic chemical compounds, metal ore processing plants, and electroplanting plants. Depending on the type of processing plant, it contains various types of pollutants, including biogenic elements, acids,

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alkali, salts and heavy metals. Toxic inorganic sewage may contain: heavy metals, inorganic acids, alkali, cyanides, etc. as the main pollutants. Such sewage is formed in metal ore and sulphur mines, in heat power plants, smelters, electroplanting plants, mineral acid works, and in explosive factories.

3.2. Biotransformation of Solid Waste in Organic Wastewater The presence of SRB has been noted in practically all reactors treating various types of organic sewage, e.g. dairy sewage (Baena, 1998, 1999, 2000; Hernandez, 2000), however they still play a small role due to the low concentration of sulphates ca. 258 g/m3 (Talik, Kutera, 1997). The source of sulphates for SRB may be not only easily soluble compounds such as Na2SO4, but also insoluble mineral phases such as hanebachite (CaSO4), jarosite (KFe3[OH]6 (SO4)2), anglesite (PbSO4) or barite (BaSO4) (Karnachuk et al., 2003). A good source of sulphate ions are solid anthropogenic wastes such as phosphogypsum formed during the production of phosphoric acid. Addition of sulphates to primarily non-sulphur sewage results in high-sulphur sewage similar to sewage produced during the manufacture of molasses (2.9 g SO4/L), citric acid from sugar cane (2.5–4.5 g SO4/L) or wood industry (1–2 g SO4/L) (Colleran et al., 1995). High concentration of sulphates in sewage hampers treatment using methanogenesis; in turn they can be purified with application of sulphidogenesis. In reactors treating high-sulphur sewage, SRB are solely responsible for the final stages of biodegradation of organic pollutants (Colleran et al., 1995). Biodegradation of two different industrial wastes including sewage and solid waste seems both an interesting and indispensable procedure for economic reasons. Biotransformation of phosphogypsum in various industrial sewages such as refinery-industrial sewage, dairy sewage or distillery decoctions has already been described (Wolicka & Kowalski, 2006a; 2006b; Wolicka, 2008b). Products of phosphogypsum biotransformation and biodegradation of organic compounds in stationary cultures were carbonates and/or phosphates. The obtained results indicate the possibility of obtaining secondary post-culture deposits that can later be applied e.g. as fertilizers in farming. Due to the lack of hydrolytic enzymes in most SRB, application of a two-step process of the treatment of organic sewage with phosphogypsum seems a good solution (Kaufman et al., 1996; Deswaef et al., 1996) (Figure 2). In the first phase, the biological reactor is colonized by an assemblage of acidogenic and sulphate reducing bacteria. Acidogenic bacteria are capable of producing acetate from organic compounds that are not easily accessible to SRB. In turn, the produced acetate can be a good energy substrate for the second group of bacteria. In the second stage, the reactor is colonized mainly by SRB that use simple organic compounds flowing from the first reactor. In both reactors waste gypsum is the source of sulphates. This method of sulphidogenesis allows for effective phosphogypsum biodegradation in the organic sewage setting, which can at first contain organic compounds that are not easily accessible for SRB (Deswaef et al., 1996). The process described herein can be conducted in bioreactors basing on the flooded deposit with biofilm. A crucial problem in these methods is the possibility of biofilm overgrowth on grains of filling material, a process that can be additionally accelerated by sparingly soluble gypsum. The possible solution may be application of reactors in which the circulating elements are the filling

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material grains with biofilms. As a result, excessive biofilm is removed and good contact of microorganisms with the solution of sewage with phosphogypsum is secured. Due to continuous circulation, the accumulation of non-reduced calcium sulphate in the biofilm is simultaneously hampered.

Figure 2. Treatment of organic waste water with phosphogypsum in two stages process.

The rule of biological deposits is based on removal of pollution from sewage by microorganisms in the biofilm deposited on various types of carriers. In biological sewage treatment many types of biological deposits are applied. Selection of the deposit type depends on technological parameters of the sewage treatment plant, type of sewage and degree of sewage treatment. The most commonly applied deposits are biological sprinkling beds comprising a container with various types of rough surface fillers e.g. pebbles, slag, plastic elements, etc. The carrier surface is covered by biofilm composed of microorganisms such as bacteria, fungi or protozoans. During the process, selection and adaptation of microorganisms to respective treatment conditions takes place. The biofilm microorganisms utilize organic and some inorganic compounds from sewage and cause decrease of their content in the sewage. Effective deposit activity depends on various parameters such as content and composition of treated sewage, temperature, pH. Biological deposits can be used in the treatment of many different types of sewage. Characteristics of the biofilm are based on observation of microorganisms comprising it. It is also tested whether biofilm ripped off the surface of fillers does not occur in the runoff of the incubating device. Two main types of bioreactors are currently applied in anaerobic sewage treatment: Reactors with biofilm on solid material (filler), such as plastic, ceramic and glass particles, sand grains, expanded clay pellets, etc. This type pf bioreactor is characterized by the fact that biomass is retained in the reactors and sewage can flow faster than it is assumed from the growth speed of microorganisms.

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Reactors without solid phase, in which granules composed of anaerobic microorganism assemblages are formed naturally. This bioreactor type increases the settling of microorganisms, what is aided by the separation of biomass from the treated sewage. Application of settling tanks in such bioreactors allows initial concentration of biomass prior to its collection, and at the same time decreases the content of microorganisms in the discharged sewage. Application of multi-stage continuous cultures in the biotransformation of phosphogypsum seems very interesting due to the occurrence of rare earth elements. Recovery of these elements using sulphidogenesis seems purposeful due to the possibility of accumulation in the biofilm of the deposit, what has been well documented (Mueller & Steiner, 1992; Kowalski et al., 2002). The problem of accumulation of heavy metals in the secondary excessive deposit is very crucial, because on the one hand it hampers wide usage of the deposits e.g. in agriculture, but on the other hand, seems a good method of recovery of the REE scattered in phosphogypsum. It has been determined (Kowalski, 2002) that REE percentage content is higher in secondary post-culture deposits that in phosphogypsum. Due to the fact that in subsequent treatment stages the deposit mass and the content of lanthanoids decrease and assuming that the system is closed, it can be stated that the REE are retained on the biofilm (Kowalski, 2002). Problem of the concentration of scattered metals in various industrial wastes is not only the case of phosphogypsum, but is also crucial in the treatment of other solid inorganic wastes. The effect of this process are sulphides of various metals, which can be recovered also using biological methods, although in aerobic conditions, e.g. using chemolithoautotrophic sulphur bacteria.

3.3. Removal of Heavy Metals from Inorganic Wastewater Heavy metals are metallic elements with density over 4.5 g/cm3 and atomic mass exceeding 50. This group comprises such metals as: arsenic, manganese, zinc, chrome, iron, cadmium, lead, nickel, copper, mercury, cobalt, and molybdenum. Their characteristic feature is the ability of very long persistence in the natural environment, what is linked with sparing solubility of some heavy metals chemical compounds and also with large ability to accumulate in living organisms. Heavy metals can be subdivided into four groups: 1. elements with very high potential hazard to environment, e.g.: cadmium, mercury, chrome, silver, zinc, gold, antimony, tin, thallium; 2. elements with high potential hazard to environment, e.g.: molybdenum, manganese, iron; 3. elements with medium potential hazard to environment, such as vanadium, nickel, cobalt and wolfram; 4. elements with low potential hazard to environment, e.g. zirconium, tantalum, lanthanum, niobium.

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magmatic and post-magmatic processes; metamorphic processes (transformation of sedimentary or magmatic rocks in high temperature or pressure, etc.); hypergenic processes (taking place under hydro-, atmo- and biospheric factors, e.g. weathering, erosion, transportation, sedimentation, etc.).

Due to these processes the primary occurrence of heavy metals is transformed. Migration of heavy metals as well as their uncontrolled appearance in the environment is greatly influenced by anthropogenic processes. Industrial activity, e.g. exploitation of metal ores, transport of ore to processing plant, transformation, final management and utilization of used products are the main anthropogenic sources of heavy metals. The resulting solid and gaseous wastes as well as sewage are accumulated in water, soil and atmosphere. High content of heavy metals can be found in sewage deposits and industrial sewage. Industrial waste containing heavy metals derive mainly from: smelters, electroplating plants, tanning, fertilizer, pesticide, dyeworks, textile, and electrochemical industries, from plants producing batteries, accumulators, catalysts, etc. Copper in sewage is derived from e.g. metallurgy, dyeworks, and textile industries, and is also emitted during production of pesticides and fertilizers. Electroplating and paper industries, refineries, and steelworks supply high content of nickel to the environment, whereas production of batteries and paints, textile and plastic industries, polymer stabilizer industry, printing and graphic plants deliver high contents of zinc. Heavy metals are also introduced to water with industrial sewage and wastes, with effluents from fields or smelter dumps. Products of sewage treatment are sewage deposits in which the heavy metal concentration distinctly exceeds that in sewage. The harmful influence of heavy metals on living organisms and the natural environment is undisputable. Therefore actions should be undertaken to eliminate or/and restrict their emission, as well as minimize their negative influence. Industrial plants in which sewage with heavy metals are produced should have installations for their preliminary treatment before sewage is sent to the receiver. Treatment of isolated sewage streams is more effective that treatment of mixed sewage. Sewage containing heavy metals are commonly treated using: chemical (neutralization, reduction and/or oxidation, precipitation), physical and chemical (sorption, extraction, ion exchange) and electrochemical methods. The correct method depends on the type of sewage, their content, phase and concentration of particular particles and required treatment degree. Recently, biological methods are also more frequently applied (Figure 3); these methods allow recovery of heavy metals and cause the transformation of toxic cations of heavy metals into sparingly soluble sulphides, a desired effect particularly in solid waste management (Ekstrom et al., 2008; Gibert et al., 2002).

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Figure 3. Biological methods to the remove heavy metals from waste water.

Bacterial reduction of sulphates allows simultaneous removal of metal sulphates by transforming them into sparingly soluble sulphides, biodegradation of organic compounds and lower acidification of the environment. Data on the influence of heavy metals on SRB are not univocal. According to some authors, even low concentrations of heavy metals can hamper SRB activity, whereas according to others, these microorganisms have low sensitivity to the presence of heavy metals because sulphide ions produced by them during sulphate reduction allow their precipitation into nonsoluble and non-toxic deposits. These discrepancies may come from the fact that the cultures are variously multiplied, i.e. at different composition, pH, and content of introduced microorganisms as well as the initial form of the introduced metal (Koschorreck, 2008). Application of SRB in bioremediation of sewage containing high concentration of heavy metals requires knowledge on the toxicity of these metals (Dvorak et al., 1992; Hass & Polpraset, 1993; Karnachuk et al., 2003). Heavy metals play also a crucial role in the metabolism of microorganisms, e.g. bacteria taking part in the regulation of biochemical processes, stabilization of cell structures or catalysis of enzymatic reactions. SRB can be thus be utilized in treatment of sewage and mine water from toxic metals, and the microorganisms applied in this process should effectively reduce sulphates, tolerate changes in pH and be resistant against toxic metals. Microorganisms conducting dissimilative sulphate reduction contribute to the treatment of mine and metallurgy sewage, as well as effluents from municipal and industrial waste dumps. Application of SRB in the demobilization of various heavy metals has been described in a number of papers, e.g. the influence of chromium, nickel, manganese, copper, and zinc on Desulfovibrio vulgaris and Desulfovibrio sp. has been described by Cabrera et al. (2006). Removal of heavy metals in short-term bench scale upflow anaerobic packed bed reactor was described by Jong & Parry (2003), and the influence of copper and zinc on the mixed SRB population was studied by Utgikar et al. (2003). Rafida (2008) notes the significant role o f biofilm formed during SRB activity.

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3.4. Co-Existence and Competitions between of the Different Groups of Microorganisms under Anaerobic Conditions Anaerobic organic sewage treatment is based mainly on the activity of several groups of bacteria: fermentation, acetogens, methanogens, sulphate reducing bacteria and denitrifying bacteria. Some role is played also by bacteria reducing iron and oxygenated forms of other metals, but the content of these bacteria is very low. The most well known and commonly applied method of anaerobic sewage treatment uses the activity of the first three mentioned microorganism groups. Each microorganism group partly oxygenates an organic compound to the relevant end products, which are next assimilated by the next link of the food chain till complete oxygenation (according to the scheme in Figure 4). This system of biological sewage treatment in the biodegradation of organic pollutants begins with the activity of fermentation bacteria, which are responsible for hydrolysis and fermentation of particular organic compounds representing the main pollutant in the sewage. Many microorganisms display their ability to fermentation, in course of which various organic compounds are formed becoming an easily accessible carbon source for SRB. The product of acetic acid fermentation, carried out by acetogens, is acetic acid. This reaction is catalyzed by enzymes produced mainly by acetate bacteria, but takes place also during other biochemical transformations. CH3CH2OH + O2 → CH3COOH + H2O + 490 kJ (118 kcal) Various acetate bacteria species display smaller or larger abilities of further oxidation of acetic acid according to the following formula: CH3COOH + 2O2 → 2CO2 + 2H2O

Figure 4. Possible stages of organic compounds biodegradation by SRB and methanogens. MA – methanogenic Archea, SRB – sulphate reducing bacteria.

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The ability of lactic acid fermentation occurs in bacteria of the proper lactic acid fermentation. These bacteria oxidize simple carbohydrates and disaccharides to lactic acid (as the main product) and to various by-products. Lactic acid fermentation was observed e.g. in Micrococcus, Escherichia and Microbacterim. This fermentation can be sub-divided into two types: homofermentation – where lactic acid is the main product and there is a low content of by-products; and heterofermentation – where more by-products are formed in relation to lactic acid. Lactic acid fermentation is a typical case of metabiosis i.e. growth of one group of microorganisms after another. In the first stage heterofermentation bacteria develop that acidify the environment and create favourable conditions for the development of bacteria from proper lactic acid fermentation (homofermentation). Ethanol is also one of the byproducts of homofermentation, resulting from the anaerobic decomposition of carbohydrates into ethylic alcohol and carbon dioxide. The products of butyric acid fermentation may be acetic acid, succinic acid and ethanol. This fermentation is carried out by Clostridium (C. acetobutylicum, C. butylicum), during which butanol is formed instead of butyric acid, and acetone instead of acetic acid. At first this fermentation is identical with butyric acid fermentation. However, when the produced acids decrease the reaction to about 4, fermentation changes course and neutral compounds such as butanol and acetone are formed instead of acids. There are also bacteria species which reduce acetone to isopropyl alcohol. It is worth noting that all fermentation products are an easily accessible carbon source for SRB. The last stage of biological sewage treatment is the activity of methanogens. Using hydrogen, carbon dioxide and acetate produced by acetogens, they produce methane. It is estimated that about 1/3 methane produced is the product of CO2 reduction, and the rest – of acetate decarboxylation. The two main advantages of this treatment method is the formation of low quantities of excessive sludge, difficult to utilize, and the possibility of using the formed biogas as fuel. The energetic value of biogas increases with the content of methane, but it should be remembered that the content of methane in biogas is inversely proportional to the concentration of SO42- in the treated sewage. Sulphates that are present in variable content in organic sewage do not directly influence the activity of methanogens, but they favour the selection of SRB, which according to reaction stoichiometry transform SO42- to H2S. H2S has toxic influence on methanogens, and decreases the quality of biogas. SRB have been noted in all types of bioreactors purifying sewage in anaerobic conditions. In such conditions they compete with various bacteria groups for the available organic compounds at almost all decomposition stages except hydrolysis, because most microorganisms capable of dissimilative sulphate reduction do not develop hydrolytic enzymes. The only exception is Archeoglobus fulgidus. In reactors treating high-sulphur sewage e.g. from pulp and paper, textile, pharmaceutical, metallurgic, paint and varnish, and plastics industries SRB are solely responsible for the final stages of organic pollutants decomposition (Colleran et al., 1995). Some SRB species such as Desulfuromonas acetoxidans may form syntrophic relationships with photosynthesizing bacteria belonging to green sulphur bacteria (Chlorobiaceae). Under sun light the phototrophs assimilate carbon dioxide and oxidize hydrogen sulphide, which is the electron donor, to elemental sulphur that is released from the cell. In the following stage Desulfuromonas reduces it to hydrogen sulphide, with

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simultaneous oxidation of acetate and production of carbon dioxide. In some cases syntrophic relationships have been noted in other SRB species, e.g. Desulfovibrio. The common development of these bacteria is an example of a syntrophic relationship, in which substrates are distributed in two directions The product of the activity of one bacteria group becomes the substrate for the next group (Figure 5). As mentioned above, SRB co-occur in all types of bioreactors applied in anaerobic treatment of sewage along with methanogens. The growth and increase of SRB activity depends on winning the rivalry with other bacteria groups on each stage of biodegradation. Competition of various microorganism groups in the presence of sulphates may take place in several different stages of the biodegradation process (Colleran et al., 1995): in the first stage of biodegradation competition with fermentation bacteria for products of biodegradation of polymeric compounds i.e. for simple monomeric compounds; competition with acetogens producing hydrogen for indirect fermentation products such as propionate or ethanol, which are an optimal carbon source for most SRB; competition with homoacetogens producing acetate for hydrogen, and at the end of biodegradation with methanogens for hydrogen or acetate.

Figure 5. Sulphur cycle in a syntrophic relationship of a Desulforomonas/Desulfovibrio culture with green sulphur bacteria.

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During sewage treatment conditions can be created to favour the development of a specific microorganism group. The dominating group of microorganisms depends on the concentration of sulphates in sewage and the chemical oxygen demand (COD). The calculated COD/SO4 ratio in sewage supplies information about the possible selection of a specific group of microorganisms (Hao et al., 1996). If the value equals 0.67 or is lower, SRB dominate in the system and the decomposition end-products is not a mixture of CO2 and CH4, but CO2 and H2S. It seems that the co-existence of the two groups of microorganisms is possible in the range of COD/SO4 from 1.7 to 2.7; below 1.7 SRB prevail (Clancy et al., 1992). If the ratio is higher, then more organic compounds are biodegraded during methanogenesis. It is estimated that at COD/SO4 > 10 sulphides are not produced (Oude Elferinck et al., 1998). With known concentration of both components, i.e. TOC and S-SO4 precise conditions for the development of a particular microorganism group can be defined and repeated (Oude Elferinck et al., 1998). Decomposition of organic matter with the use of SRB is becoming to be applied in the treatment of high-sulphur waste. In comparison with the method based on methanogenesis, its advantage is the shorter duration of biodegradation, precipitation of heavy metals into poorly soluble sulphides and elimination of pathogenic organisms (toxic activity of hydrogen sulphide). The need to oxygenate hydrogen sulphide to elemental sulphur and slightly higher volume of excessive sludge than in methanogenesis is on the other hand a disadvantage.

3.5. Biological Methods of Sulphide Treatment – By-Products of Sulphate Reduction (Photosynthesizing and Chemoautotrophic Bacteria) The process of anaerobic sewage treatment using sulphidogenesis is linked with the necessity to remove hydrogen sulphide produced by SRB. This requirement seems the most important disadvantage of the method; therefore many reports regarding this topic have been published (Kobayashi et al., 1983; Khanna et al., 1996; Lee & Kim, 1998; Tichy et al., 1998). So far, the often proposed solution is chemical (abiotic) oxidation of hydrogen sulphide to elemental sulphur; this process is, however, very expensive due to the application of catalysts and energy demanding aeration. Therefore, solutions with application of microorganisms capable of replacing chemical removal of hydrogen sulphide are sought for. It seems that the most appropriate microorganisms are photosynthesizing and chemolithoautotrophic bacteria. Both groups comprise microorganisms capable to oxidation of hydrogen sulphide, although the metabolism linked with this process is entirely different. The most important problem during application of microorganisms in removing hydrogen sulphide is the selection of tribes and physical-chemical conditions producing elemental sulphur in course of oxidation; the sulphur would then be emitted from the reactor. It is thus indispensable that the final effect of sulphur transformation in anaerobic high-sulphur sewage treatment using sulphidogenesis and additional refinery stages would be elemental sulphur being a potential source in many industrial branches. It poses much less hazard to the natural environment during stacking than other solid waste containing sulphur, e.g. waste gypsum.

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3.5.1. Microorganisms applied in biological hydrogen sulphide treatment. Photosynthesizing sulphur bacteria. This is the largest and probably the most appropriate group with regard to application in hydrogen sulphide removal that encompasses microorganisms capable of oxidizing hydrogen sulphide to elemental sulphur (Gemerden, 1986; Eraso & Kaplan, 2001). It comprises a diverse group of microorganisms classified into many taxonomic units which share the ability to carry out processes depending on presence of light. Typically, four sub-groups are distinguished: green sulphur bacteria, purple sulphur bacteria, purple non-sulphur bacteria and the Chloroflexus group comprising non-sulphur green bacteria. Due to their ability to cumulate elemental sulphur the most appropriate for application in biotechnology of hydrogen sulphide oxidation are the first two groups. Green sulphur bacteria conduct photosynthesizing processes in which the electron sources include reduced sulphur compounds such as sulphides and tiosulphate. These bacteria contain high amounts of bacteriochlorophyll c, d, e, as well as much lower quantities of bacteriochlorophyll a (Brock & Madigan, 2006; Baneras et al., 1999). They are assigned to the order Chlorobiales, which comprises such species as Chlorobium limicola, Chlorobium limicola thiosulfatophilum, and Pelodictyon sp. They occur in aqueous environs, particularly in thermally stratified lakes, but can also be found in saline conditions or in highly exceeded temperatures. Many species of green bacteria are easily isolated from the natural environment and multiplied in laboratory conditions, what allows potential application in biotechnology. The most important feature of the group with regard to this fact is the ability to cumulate elemental sulphur outside the cells, what distinguishes them from another group of photosynthesising bacteria – purple sulphur bacteria of potential significance in the process. Microorganisms belonging to purple sulphur bacteria contain high amounts of bacteriochlorophyll a and have the ability to cumulate sulphur, but only within the cell, with the exception of the family Ectotiorhodospiraceae that cumulates sulphur outsides the cell (Eraso & Kaplan, 2001). The basic source of electrons in the photosynthesizing processes includes hydrogen and hydrogen sulphide, and the stored sulphur may potentially be oxidized to sulphate. The bacteria belong to the order Chromatiales, and the best known species is Chromatium okeni. Chemolithoautotrophic sulphur bacteria. The second large group potentially applicable in biotechnology of hydrogen sulphide removal comprises autotrophic bacteria carrying out oxidation processes of reduced inorganic compounds such as sulphides, hydrogen sulphide, iron (II) as well as tiosulphate and elemental sulphur. The processes supply to the bacteria cells energy needed to bind carbon dioxide. The group includes such species as Acidithiobacillus tiooxidans, A. ferrooxidans, Thiobacillus thioparus, and Starkeya novella. Chemolithoautotrophic sulphur bacteria are generally applied in processes of bioleaching of metals from low-percentage ores or other waste containing metal sulphides, as well as in bioremediation (Third et al., 2000; Suzuki, 2001; Pacholewska, 2004; Suzuki & Suko, 2006). Due to this fact these bacteria could theoretically be appropriate to apply in recovery of metals from the remaining deposit, e.g. after utilization of phosphogypsum during sulphidogenesis. Such deposit, due to heavy metal content in phosphogypsum would have a high amount of insoluble metal sulphides that potentially could be utilized. However, application of chemolithoautotrophic bacteria in refinement of sewage rich in hydrogen sulphide has created a number of problems. First of all, despite the existence of species

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producing sulphur such as Thiobacillus thioparus, a large number of bacteria species oxidize hydrogen sulphide directly to sulphates. Secondly, these bacteria often require lower pH in the environment, and thirdly, the effluent after refinement using sulphidogenesis may contain organic compounds distinctly hampering the activity of chemolithoautotrophic sulphur bacteria. Due to this fact, the present application of these bacteria is restricted to attempts of metal bioleaching and in refinement of mineral sewage containing e.g. tiosulphates.

3.5.2. Application of photosynthesizing sulphur bacteria in removal of hydrogen sulphide from sewage Attempts to apply green and purple sulphur bacteria in the removal of hydrogen sulphide are justified in all processes that generate the formation of sewage strongly contaminated by sulphides, hydrogen sulphide and tiosulphates. Such sewage is formed during gas desulphurization, in crude oil processing plants, chemical plants, and mostly during treatment of high-sulphur sewage using aerobic methods. Solutions proposed hitherto are based on the utilization of photosynthesizing bacteria populations in photoreactors, in which are ensured conditions favourable for their growth. The most important and practically the only necessity is the assurance of a light source, what is linked with some cost, but due to the possibility of using day-light, the cost is distinctly lower than in the case of a chemical process. Lee & Kim (1998) proposed the application of the so-called optical-fibre bioreactor, to which light is brought by optical fibres. The bacteria Chlorobium limicola thiosulfatophilum were used in the bioreactor with successful results. Henshaw et al. (1998) also applied Chlorobium limicola in a system similar to a chemostat, in a suspended-growth continuous stirred tank reactor, which was illuminated by a lamp emitting infrared light. The process was highly effective, with 90% conversion of sulphides in the solution to elemental sulphur. Henshaw et al. (1999) tested the influence of the material used to construct the bioreactor (mainly various types of plastics permitting infrared radiation) and showed the lack of significant influence of the applied material on the growth of Chlorobium limicola. It is worth noting that a rather low light intensity was applied in this experiment, from 3.4 ∙ 10-3 to 4.7 ∙ 10-3 W m-2. Application of photosynthesizing bacteria is linked with the important issue of organic compounds presence in sewage flowing to the reactor. Due to anaerobic conditions SRB may grow in the reactor; at this stage of refinement this is highly undesirable, influencing the effective reduction of hydrogen sulphide pollution. Theoretically, the presence of organic compounds may be unfavourable to the metabolism of photosynthesizing bacteria, although it seems that they are partly capable to utilize simple organic compounds. This problem is crucial during attempts of applying photosynthesizing bacteria to remove hydrogen sulphide from sewage earlier treated during sulphidogenesis. Such sewage may contain also indistinct amounts of organic matter and may be also the source of SRB. Borkowski & Wolicka (2007a) applied an assemblage of photosynthesizing bacteria isolated from the natural environment on a flooded deposit and indicated that yeast extract flowing with synthetic medium may distinctly inhibit the efficiency of sulphide oxidation. In turn, in Borkowski & Wolicka (2007b) the synthetic medium flowing onto the flooded deposit colonized by photosynthesizing bacteria was replaced by filtered influent from a sulphidogenic bioreactor, in which phosphogypsum with distillery decoctions was treated. In this case, 60% effectiveness of sulphide content reduction was obtained (from 163 mg L-1 to 70 mg L-1) with partial conversion to elemental sulphur.

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Figure 6. Biological oxidation of hydrogen sulphide after sulphidogenesis.

During simultaneous treatment of organic sewage and solid gypsum waste such as phosphogypsum using anaerobic treatment by sulphidogenesis, it is possible to refine the sewage already after the main process (Figure 6). The process should comprise the following stages: 1. The effluent from the anaerobic reactor after sulphidogenesis contains sulphides and hydrogen sulphide; after primary filtering it is transported to a bioreactor with photosynthesizing bacteria. The effluent can contain also organic compounds and SRB, the number of which should distinctly decrease after filtration. The effect is almost complete reduction of the remaining organic pollutants and elemental sulphur. 2. The solid waste after phosphogypsum may contain high amounts of calcite and slightly less phosphates and due to this fact may be applied in agrotechnical activities. However, if the phosphogypsum contained a high content of heavy metals or if these metals flew with sewage, after utilization using sulphidogenesis the deposit is enriched in sulphides of heavy metals. Such deposit, beside the fact that it is very toxic, is also a significant source of metal that can be recovered during bioleaching by natural or selected communities of chemolithoautotrophic sulphur bacteria representing e.g. Thiobacillus and Acidithiobacillus.

4. CONCLUSION At present it is obvious that natural environment protection is a basic and essential task. Taking into account the great volume of sewage produced in anthropogenic processes by industrial plants, it is important to find solutions aimed at decreasing the influence of toxic and dangerous sewage on the natural environment. One of the basic methods of solving the water quality issue is rational and professional approach to problems linked with the treatment of various types of organic sewage. It is commonly known that biological methods using microorganisms can be applied in the biodegradation of hazardous petrochemical waste. Many groups of anaerobic bacteria are able to biodegrade various types of organic waste to non-toxic compounds or even to inorganic compounds after complete mineralization. Biological methods do not require introduction of chemical compounds to the environment that would negative influence the biocenosis. Most of all, biological methods take place naturally in the environment, and

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anthropogenic influence in this case is focused only on stimulating their range and intensity. Final products of microbiological decomposition include carbon dioxide and water. Simultaneous biodegradation of several industrial wastes, e.g. organic sewage and solid waste seems an interesting and optimal solution from the point of economy. Costs linked with simultaneous biodegradation of two arduous industrial wastes are almost always much lower than separate treatment processes. Application of SRB in biological treatment of sewage is becoming popular and is entirely justified.

ACKNOWLEDGMENTS I would like to thanks Dr Andrzej Borkowski for his help in preparing of this manuscript.

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Mizuno, O., Takagi, H. & Noike, T. (1998). Biological sulfate removal in an acidogenic bioreactor with an ultrafiltration membrane system. Water Science and Technology, 38, 4-5. Mueller, F. R. & Nielsen, P. H. (1996). Characterization of thermophilic consortia from two souring oil reservoirs. Applied and Environmental Microbiology, 62, 3083. Mueller, R. F. & Steiner, A. (1992). Inhibition of anaerobic digestion caused by heavy metals. Water Science and Technology, 26, 835-846. Oude Elferinck, S. J. W. H., Vorstman, W. J. C., Sopjes, A. & Stams, A. J. M. (1998). Desulforhabdus amnigenus gen.sp.nov., a sulfate reducers isolated from anaerobic granular sludge. Archives of Microbiology, 164, 119-124. Pacholewska, M. (2004). Bioleaching of galena flotation concentrate. Physicochemical Problems of Mineral Processing, 38, 281-290. Postgate, J. R. (1984). The sulphate reducing bacteria. Cambridge University Press. Przytocka-Jusiak, M., Rzeczycka, M., Ponichtera, E. & Mycielski, R. (1997). Rozkład benzenu w hodowlach termofilnych bakterii redukujących siarczany. In I cześć materiałów V Ogólnopolskiego Sympozjum Naukowo-Technicznego ―Biotechnologia Środowiskowa‖ Ustroń-Jaszowiec, 215-222. Rabus, R., Nordhaus, R., Ludwig, W. & Widdel, F. (1993). Complete oxidation of toluene under strictly anoxic conditions by a new sulfate-reducig bacterium. Applied and Environmental Microbiology, 59, 1444-1451. Rafida, A. I. (2008). Removal of heay metals from rainwater in batch reactors with sulphate reducing bacteria (SRB). Proceeding of World Academy of Science, Enegineering and Technology, 33, 23-27. Rees, G. N., Harfoot, C. G. & Sheehy, A. J. (1997). Amino acid degradation by the mesophilic sulfate-reducing bacterium Desulfobacterium vacuolatum. Archives of Microbiology, 169, 76-80. Reichenbecher, W. & Schink, B. (1997). Desulfovibrio inopinatus, sp. nov., a new sulfatereducing bacterium that degrades hydroxyhydroquinone (1,2,4-trihydroxybenzene). Archives of Microbiology, 168, 338-344. Rozanova, E. P., Borzenkov, I. A., Tarasov, A. L., Suntsova, L. A., Dong, Ch. L., Belyaev, S. S. & Ivanov, M. V. (2001). Microbiological processes in a high-temperature oil field. Microbiology, 70, 102-110. Rueter, P., Rabus, R., Wilknes, H., Aeckersberg, F., Rainey, F. A., Jannasch, H. W. & Widddel, F. (1994). Anaerobic oxidation of hydrocarbons in crude oil by new types of sulphate-reducing bacteria. Nature, 372, 455-457. So, C. M. & Young, L. Y. (1999). Initial reactions for anaerobic alkane degradation by the sulfate reducer strain AK-01. Applied and Environmental Microbiology, 65, 5532-5540. Suzuki, I. (2001). Microbial leaching of metals from sulfide minerals. Biotechnology Advances, 19, 119-132. Suzuki, Y. & Suko, T. (2006). Geomicrobiological factors that control uranium mobility in the environment: update on recent advances in the bioremediation of uraniumcontaminated sites. Journal of Mineralogical and Petrological Sciences, 101, 299-307. Szewzyk, R. & Pfenning, N. (1987). Complete oxidation of catechol by stricly anaerobic sulfate reducing Desulfobacterium catecholicum sp.nov. Archives of Microbiology, 147, 163-168.

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Third, K. A., Cord-Ruwisch, R. & Watling, H. R. (2000). The role of iron-oxidizing bacteria in stimulation or inhibition of chalcopyrite bioleaching. Hydrometallurgy, 57, 225-233. Tichy, R., Janssen, A., Grotenhuis, J. T. C., van Abswoude, R. & Lettinga, G. (1998). Oxidation of biologically-produced sulphur in a continuous mixed-suspension reactor. Wat. Res., 32(3), 701-710. Utgikar, V. P., Tabak, H. H., Haines, J. R. & Govind, R. (2003). Quantification of toxic and inhibitory impact of copper and zinc on mixed cultures of sulfate-reducing bacteria. Biotechnology and Bioengineering, 82, 306-312. Voordouw, G., Armstrong, S. M., Reimer, M. F., Fouts, B., Telang, A. J., Shen, Y. & Gevertz, D. (1996). Characterization of 16S rRNA genes from oil field microbial communities indicates the presence of a variety of sulfate-reducing, fermentative, and sulfide-oxidizing bacteria. Applied and Environmental Microbiology, 62, 1623-1629. Warren, E., Bekins, B. A. & Godsy, M. (1999). Inhibition of acetoclastic methanogenesis by crude oil from Bemidji, Minnesota. Widdel, F. & Bak, F. (1992). Gram-negative mesophilic sulfate-reducing bacteria. In A. Balows, H. G. Truper, M. Dworkin, W. Harder, & K. H. Schleifer (Eds.), The prokaryotes, 2nd end, chapter 12. Wilkes, H., Boreham, C., Harms, G., Zengler, K. & Rabus, R. (1999). Anaerobic degradation and carbon isotopic fractionation of alkylobenzenes in crude oil by sulphate-reducing bacteria. Organic Geochemistry, 31, 101-115. Wit, R. (1992). Sulfide containing environments. In J. Ledeberg (Eds.), Encyclopedia of Microbiology, Academic Press, New York, 105-121 Wolicka, D. (2006). Biotransformation of phosphogypsum in cultures of bacteria isolated from petroleum. In Proceedings of the International Conference Protection and Restoration of the Environmental VIII, Create, 2006. Nikas Books, 217-218. Wolicka, D. (2008a). Biodegradation-the natura metod for liquidation enironmental polluted by petroleum products. Oil and Gas Institute, Prace, 150, 675-680. Wolicka, D. (2008b). Biotransformation of phosphogypsum in wastewaters from dairy industry. Bioresource Technology, 99, 5666-5672. Wolicka, D. & Borkowski, A. (2007a). Participation of sulfate-reducing bacteria in biodegradation of organic matter in soils contaminated with petroleum products. Archives of Environmental Protection, 33, 93-99. Wolicka, D. & Borkowski, A. (2007b). Activity of a sulphate reducing bacteria community isolated from an acidic lake. EANA 07, 7th European Workshop on Astrobiology, Turku, Finland, October 22-24, 2007, 98. Wolicka, D. & Borkowski, A. (2008). Oil-derived products as the source of sulphate-reducing bacteria biotransforming phosphogypsum. Polish Journal of Environmental Studies, 17, 592-595. Wolicka, D. & Borkowski, A. (2009). Phosphogypsum biotransformation in cultures of sulphate reducing bacteria in whey. International Biodeterioration & Biodegradation, 63, 322-327. Wolicka, D. & Kowalski, W. (2005). Utilization of different carbon compounds by sulphatereducing bacteria In medium with phosphogypsum. Archives of Environmental Protection, 31, 105-112. Wolicka, D. & Kowalski, W. (2006a). Biotransformation of phosphogypsum on distillery decoctions (preliminary results). Polish Journal of Microbiology, 55, 147-151.

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Wolicka, D. & Kowalski, W. (2006b). Biotransformation of phosphogypsum in petroleumrefining wastewaters. Polish Journal of Environmental Studies, 15, 355-360. Wolicka, D., Kowalski, W. & Boszczyk-Maleszak, H. (2005). Biotransformation of phosphogypsum by bacteria isolated from petroleum-refining wastewaters. Polish Journal of Microbiology, 2, 54. Zhang, W. & Bouwer, E. J. (1997). Biodegradation of benzene, toluene and naphtalene in soil–water slurry microcosms. Biodegradation, 8, 167-175.

In: Fluid Waste Disposal Editor: Kay W. Canton, pp. 81-112

ISBN: 978-1-60741-915-0 © 2010 Nova Science Publishers, Inc.

Chapter 3

UTILIZATION OF WATER AND WASTEWATER SLUDGE FOR PRODUCTION OF LIGHTWEIGHT-STABILIZED CERAMSITE 1

Zou Jinlong 1, 2, Yu Xiujuan 1, Dai Ying 3 and Xu Guoren 2 Department of Environmental Science, School of Chemistry and Materials Science, Heilongjiang University, 150080, Harbin, China. 2 State Key Laboratory of Urban Water Resource and Environment, Harbin Institute of Technology, Harbin 150090, China. 3 School of Civil Engineering, Heilongjiang Institute of Technology, Harbin, 150050, China.

ABSTRACT Disposal of wastewater treatment sludge (WWTS) and drinking-water treatment sludge (DWTS) is one of the most important environmental issues nowadays. Traditional options for WWTS and DWTS management, such as landfilling, incineration, etc., are no longer acceptable because they can cause many environmental problems. Conversion of WWTS and DWTS into useful resources or materials is of great interest and must be intensely investigated. To attain this goal, WWTS and DWTS were used as components for making ceramsite. Part I: SiO2 and Al2O3 were the major acidic oxides in WWTS and DWTS, so their effect on characteristics of ceramsite was investigated. Results show that WWTS and DWTS can be utilized as resources for producing ceramsite with optimal contents of SiO2 and Al2O3 ranging from 14–26% and 22.5–45%, respectively. Bloating and crystallization in ceramsite above 900 ℃ are caused by the oxidation and volatilization of inorganic substances. Higher strength ceramsite with less Na-Ca feldspars and amorphous silica and more densified surfaces can be obtained at 18%≤Al2O3≤26% and 30%≤SiO2≤45%. Part II: Fe2O3 and CaO were the major basic oxides, so their effect on characteristics of ceramsite was also investigated. The optimal contents of Fe2O3 and CaO are in the range of 5%–8% and 2.75%–7%, respectively. Higher strength ceramsite with more complex crystalline phases and fewer pores can be obtained at 6%≤Fe2O3≤8%. Lower strength ceramsite with more pores and amorphous phases can be obtained at 5%≤CaO≤7%, which implies that excessive Ca2+ exceeds the needed ions for producing electrical neutrality of silicate networks. Part III: To

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Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren investigate stabilization of heavy metals in ceramsite, leaching tests were conducted to find out the effect of sintering temperature, pH, and oxidative condition. Results show that sintering exhibits good binding capacity for Cd, Cr, Cu, and Pb and leaching contents of heavy metals will not change above 1000 ℃. Main compounds of heavy metals are crocoite, chrome oxide, cadmium silicate, and copper oxide, which prove that stronger chemical bonds are formed between these heavy metals and the components. Leaching contents of heavy metals decrease as pH increases and increase as H2O2 concentration increases. Leaching results indicate that even subjected to rigorous leaching conditions, the crystalline structures still exhibit good chemically binding capacity for heavy metals and it is environmentally safe to use ceramsite in civil and construction fields. It is concluded from the 3 parts that utilization of WWTS and DWTS can produce high performance ceramsite, in accordance with the concept of sustainable development.

1. INTRODUCTION WWTS is a mixture of biosolid generated in the treatment of the organic substances of municipal sewage along with disease-causing pathogens and inorganic substances such as sand and metal oxides [1, 2]. It is therefore of great significance to find a proper way to dispose it to avoid secondary pollution. The generally adopted sludge disposal is landfilling, but the option takes up valuable space and may generate methane that contributes to the greenhouse effect [3]. Another commonly used method is thermal treatment [3-6] involving incineration, gasification, and pyrolysis, which can reduce the leachability of heavy metals in the obtained materials with dramatically decreased volume of sludge [5]. This method proposes an alternative waste management technology for sludge disposal, but some of the final products still have to be deposited in landfills. DWTS mainly consists of organic and inorganic compounds in solid, liquid and gaseous forms, with variable physical, chemical, and biological characteristics [7, 8]. DWTS are frequently chemically-treated and dewatered before disposal in the developed countries [7, 8]. Typical practices of disposal are landfill, recycling, reutilization [7-9], and coagulants-regeneration. Although various alternative options for disposal, regeneration and beneficial reuse of DWTS have mainly been explored in the past decade, the search for cost effective and eco-friendly disposal options has become an urgent priority due to tighter environmental regulations and declining public acceptance of other solutions in the developed countries. In the developing countries, the untreated DWTS are normally discharged into the municipal sewers which directly flow into surface water [7]. Although, this type of disposal may seem inexpensive, it is usually ecologically unfavorable. However, the traditional options for WWTS and DWTS management are no longer acceptable because they can cause many environmental problems such as atmospheric contamination, soil contamination, water contamination, etc. Alternative options need to be explored in order to solve their handling in a more environmentally sound manner. Today, some wastes with high contents of water, inorganic/organic matter, and pathogen agents have been postulated as a precursor of materials that can be successfully employed in several environmental applications [10-12]. Converting waste into useful products can alleviate the problems of disposal and offer a new reserve for depleting resources [10-20]. For example, many inorganic/organic wastes, such as steelworks slag, mining residues, slag, paper sludge, fly ash, and even municipal solid waste have been widely tested as the alternatives to produce clay, shale, fuel, glass ceramics, adsorbent, activated carbon, cement,

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etc. [14-31]. So, the conversion of WWTS and DWTS into useful resources or materials is of great interest and must be intensely investigated. Recent research has been carried out on the reuse of WWTS for production of innovative aggregates or ceramsite, to be both used as construction materials or filter media [32-34]. Much research has successfully examined the feasibility of using clay and WWTS to produce the ceramsite [34, 35]. Valuable information about the chemical speciation of heavy metals in ceramsite and their potential environmental risks is also obtained. Results show that heavy metals are properly stabilized in ceramsite and cannot be easily released into the environment again to cause secondary pollution [36, 37]. The resulting ceramsite is made with WWTS and clay. To avoid more consumption of clay and protect the earth‘s surface environment, searching for other materials to replace clay in ceramsite production needs to be encouraged to achieve the sustainable development of natural resources. DWTS is one of the best substitutes for clay because its major components are similar to those of clay. From the viewpoint of natural resource recovery and conservation, utilization of DWTS as a substitute for clay for production of ceramsite is of great interest and significance. In our studies, WWTS and DWTS, the latter being often rich in amorphous Fe or Al oxides because of Fe or Al salts used for coagulation of source water to remove turbidity and taste and to speed sedimentation [7-9], have successfully been sintered at temperatures about 1000 ℃ to produce ceramsite. (The obtained parameters for production of ceramsite are shown as follows: DWTS/WWTS=45/55, water glass/(DWTS+WWTS)=20%, sintering temperature=1000 ℃, sintering time=35min), which has the potential of cost reduction of sludge treatment (ceramsite can be sold as a commercial product) [38]. The aim of sludgeceramsite production is to form new and useful, less soluble, and more geo-chemically stable products. The significance of utilization of DWTS and WWTS as components for sludgeceramsite production are the waste recycling and re-utilization, as well as the reduction of waste volume and the total destruction of pathogenic agents and organic pollutants. This kind of conversion can revolutionize handling of such sludge types allowing them to be reused as low-cost raw materials, rather than as waste requiring costly disposal. The main concerns in this chapter are (I) whether the impact of the specific and important constituents, such as acidic oxides (SiO2 and Al2O3) and basic oxides (Fe2O3 and CaO), on characteristics of ceramsite is significant; (II) whether it is safe to use ceramsite made from sludge containing heavy metals, especially when heavy metals are not separated but stabilized in the product; and (III) whether physical and chemical changes in the environment over time will result in an unacceptable increase in heavy metal leaching. By considering the above issues, the specific objectives are established and shown as follows: Part I: The major acidic oxides, such as SiO2 and Al2O3 (Al2O3 is considered as acidic oxide), in DWTS and WWTS may strongly affect the bloating behavior and crystal formation of the ceramsite during the heat treatment process. To validate this hypothesis, the present study has been conducted (i) to utilize DWTS and WWTS for production of ceramsite; (ii) to investigate the effect of SiO2 and Al2O3 on the physical characteristics (bulk density, particle density, water adsorption, and porosity) of ceramsite, (iii) to characterize the ceramsite within the optimal content ranges of SiO2 and Al2O3 by thermal analysis, morphological structures analysis, XRD (X-ray diffraction) and compressive strength, and (iv) to analyze the sintering mechanisms, as well as to establish effective parameters for evaluation.

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Part II: It was found that the characteristics of sintered ceramsite might be strongly influenced by the basic oxides, such as Fe2O3 and CaO, which was of great importance for the sintering process to gain more crystallization phases or liquid phases [21, 38]. The required sintering temperature for production of ceramsite might be lowered by higher content of the basic oxides in the raw materials. The formation of pores and thermal bloating might be also influenced by the basic oxides, which could activate the carbon phase to decompose to release water and gases. The study in this part attempts to address this question by investigating the following: (i) to investigate the effect of Fe2O3 and CaO (basic oxides) on the characteristics of ceramsite; (ii) to characterize the ceramsite by thermal analysis (DT-TGA), XRD, morphological structures analyses, and compressive strength measurements and (iii) to analyze the sintering mechanisms, which are useful in optimizing the sintering processes for production of ceramsite. Part III: Heavy metals leachability, binding capacities, and binding strengths in ceramsite may be governed by the major parameters such as sintering temperature, and in conjunction with the external conditions (pH and oxidative condition) [14, 36, 37]. The aim of this part is to establish effective parameters for evaluation of heavy metals stabilized in ceramsite and to analyze the binding mechanisms. Thus, laboratory experiments for evaluating short-term and long-term durability of the heavy metal compounds solidified in ceramsite and theoretical analysis are thoroughly evaluated. Specific objectives of this study are to determine the effect of sintering temperature (from 900 ℃to 1100 ℃at 25 ℃intervals), pH (from 1 to 12 at 1 unit intervals), and H2O2 concentration (0.1, 0.5, 1, 1.5, 3, and 5 mol L-1) on the stabilization of heavy metals (Cd, Cr, Cu, and Pb) in ceramsite, and to demonstrate the leaching behaviors of heavy metals in it. Results will allow the fundamental understanding and quantitative description of the stabilization of heavy metals in sludge-ceramsite. Furthermore, the evaluation of the toxicity of so stabilized/solidified heavy metals in the ceramsite is helpful to optimize the ceramsite-making processes, reduce its toxicity, and meet environmental safety requirements.

2. MATERIALS AND METHODS 2.1. Materials The WWTS used in this study was obtained from the Wen–chang Wastewater Treatment Plant, Harbin, China. The dewatering of WWTS was conducted by using a belt filter press, and cationic polymeric flocculants were used for the flocculation and dewatering of the activated sludge. The sludge cake generated from the activated sludge process is approximately 1.6×105 kg d–1 in wet weight with 24% solids, which is directly landfilled. DWTS were collected from the chemical coagulation/flocculation unit of the 3rd drinkingwater treatment plant in Harbin, China. The 3rd plant uses a conventional process with aluminum sulfate [Al2(SO4)3] as the primary coagulant and a small amount of activated silicic acid with no pH adjustment. Treatment includes mechanical lattice flocculation basins, settling basins, and fast filters. Sludge and backwash water are not discharged simultaneously. The DWTS and WWTS were treated by air–dry method and were ground into sizes below 100 μm that were sufficiently fine to be mixed homogeneously. DWTS, WWTS, and water

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glass [sodium silicate–Na2O·(SiO2)x·(H2O)y] were the raw materials for production of ceramsite [20]. The components of WWTS are shown in Table 1. In order to indicate the components of DWTS are similar to those of clay, and DWTS can be tested as a substitute for clay for production of ceramsite, the components of clay and DWTS are shown in Table 2 and Table 3. The modulus of water glass used in the study was 3.2 (Water glass is an important component for the pelletization of ceramsite and the eutectic point of the mixture can be reduced by the reaction of sodium silicate and air (Na2O·nSiO2·xH2O+CO2→Na2CO3+nSiO2+xH2O ), and by the dispersion of alkali metal oxide (i.e., Na2O et al.) originating from the decomposition of Na2CO3 (Na2CO3→Na2O+CO2 ) in the heating process [35]). By simple calculation of each inorganic matter content according to the optimal parameters (DWTS/WWTS=45/55, water glass/(DWTS+WWTS)=20%, sintering temperature=1000 ℃, sintering time=35min), it can be obtained that the ratio of SiO2: Al2O3: Fe2O3: CaO in the mixture of DWTS, WWTS, and water glass for production of ceramsite is 27.2: 15.8: 6: 3.5 (SiO2 and Al2O3 are acidic oxides and Fe2O3 and CaO are basic oxides). The simulated contents (wt.%) of tested oxide (SiO2, Al2O3, Fe2O3 or CaO) are adjusted by adding the oxide or the other three oxides (Fe2O3, CaO, and SiO2 or Al2O3) to the raw materials. All the used oxides (SiO2, Al2O3, Fe2O3 and CaO) with particle sizes below 10 μm were of the highest purity and of analytical grade. Table 1. Component analyses of dried WWTS (wt.%). component analyses SiO2

Al2O3

Fe2O3

CaO

MgO

P2O5

K2O

Others

Carbonaceous Matter

16.28

6.35

5.15

4.10

1.67

1.65

1.12

<0.89

<62.90

element analyses Zn

Fe

Mn

Si

Cu

Ca

Mg

O

Cr

0.13

3.81

0.14

8.26

0.03

3.18

1.02

48.83

0.03

P

Na

K

Al

Ni

C

S

N

Others

0.84

0.34

0.95

3.36

0.10

27.41

0.04

1.03

<0.50

Table 2. Component analyses of dried clay (wt.%). component analyses SiO2

Al2O3

Fe2O3

TiO2

K2O

MgO

CaO

Others

64.89

24.95

2.50

1.20

1.20

0.50

0.50

<0.5

Carbonaceous Matter <3.60

elements analyse Zn

Fe

Mn

Si

Cu

Ca

Mg

O

Cr

0.06

1.83

0.08

32.62

0.02

0.37

0.31

44.56

P

Na

K

Al

Ti

C

Ba

Zr

Others

0. 11

0.26

0.98

15.81

0.65

1.88

0.03

0.01

<0.40

0.02

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Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren Table 3. Component analyses of dried DWTS (wt.%).

SiO2 40.61

Al2O3 27.36

Fe2O3 6.99

CaO 2.62

Zn 0.02 Cl 0.11

Fe 4.47 Na 0.74

Mn 0.57 K 1.03

Si 20.55 Al 14.16

component analyses MgO K2O Na2O 1.89 1.28 1.05 element analyses Zr Ca Mg 0.02 1.95 1.10 Rb C Ti 0.01 6.52 0.16

Others <1.0

Carbonaceous Matter <17.20

Sr 0.02 O 48.06

Cr 0.01 Others <0.50

Table 4. Average content of heavy metals in WWTS at different wastewater treatment plants (mg kg–1). Site (China) Beijing Shanghai Shenyang Changsha Wuhan Harbin

Cd 42.20 0.19–5.55 11.82 8.1 0.70 5.9–13.5

Cr6+ 179.6 1.13–70 147.8 436.5 72–78.6 89–326.3

Cu 300 101–426 274.53 331.3 148.4 93–333.2

Pb 937.40 59.5–129 306.82 151.8 55.1 59.5–134

Table 5. Content of heavy metals in reference WWTS samples (mg kg–1). NO Minimum Medium Maximum

Cd 1 25 50

Cr 100 500 1000

Cu 100 250 500

Pb 50 500 1000

In leaching tests, the reference sludge samples were made with heavy metals by adding metal solutions [Cd(NO3)2, K2CrO4, Pb(NO3)2, and CuSO4 were of the analytical grade] into the dried WWTS, mixing and allowing them to react for 30 days. The contents of Cd, Cr, Cu and Pb were designed according to the basic data obtained through analysis of activated sludge at different places in China as shown in Table 4. The synthetic metal solution was prepared by dissolving 0.05 g L–1 of Cd2+, 0.1 g L–1 of Cr6+, 0.1 g L–1 of Pb2+ and 0.5 g L–1 of Cu2+ in deionized water. The simulated content of heavy metals was prepared by adding the tested heavy metal compounds into WWTS. The content of heavy metals added to the WWTS are shown in Table 5. The ceramsite prepared for leaching was made with this WWTS.

2.2. Methods The WWTS containing heavy metals and DWTS was treated by air-dry method and was ground into sizes below 100 μm that are sufficiently fine to be homogeneously mixed. The ceramsite for determination of the stabilization of Cd, Cr, Cu and Pb was

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made of DWTS, WWTS, and sodium silicate. The raw materials were mixed and pelletized to particle sizes of 5–8 mm and left at a room-temperature of about 20 ℃ for a few days (about 3 days) and then the samples were dried at 110 ℃ in a DHG-9070A blast roaster (China) for 24h. The heating of samples started at 20 ℃, heated at a rate of 8 ℃/min in a SX 2 -10-12 muffle furnace (China), and the samples were soaked at 200 ℃, 600 ℃, and 800 ℃ for a duration of 10min and at 1000 ℃ for a duration of 35min, and then these samples were naturally cooled until they reached room temperature . Bulk density which includes all voids and spaces in the volume, particle density which is also the apparent specific gravity of the aggregates includes all intraparticle voids, water absorption determined from the weight differences between the sintered and water saturated samples (immersed in water for 1h), and porosity ((1–bulk density/particle density) ×100%) were analyzed [34]. To achieve statistical soundness, at least three replicates were carried out for each sample. Thermal behaviors of samples were examined by thermodifferential and thermogravimetric analyses (DT–TG) using a ZRY–2P simultaneous DT–TG analyzer (China) while the samples were heated at a rate of 8 ℃/min from 20 ℃to 1080 ℃ in air. Samples weighed from 4 to 10 mg in mass, and they were put into a Pt–Rh crucible with 20 taps. All curves were evaluated using the TA–instruments software. The second derivative differential thermal curve was used for determination of peak temperature. Scanning electron microscope (SEM) and energy dispersive spectrum (EDS) analyses were conducted by using S–570 scanning electron microscope and TN–5502 X–ray energy dispersive spectrum (Japan). Compressive strength of ceramsite was analyzed by using an INSTRON 5569 automatic material testing machine (USA) while the sintered ceramsite with diameter of 6–8 mm was placed vertically on the platform of the press and was pressed at a crosshead speed of 0.5 mm min –1 until it was crushed. Toxicity of ceramsite samples was determined by using a revised method derived from toxicity characteristic leaching procedure (TCLP), a standard method used to determine waste leaching toxicity, updated on the basis of hazardous waste extraction procedure (EP) by USEPA [36]. By using this method, the leaching test was conducted with the solution prepared at a liquid–solid ratio of 1L/200g, and stirred at 110rpm for 24h or 30d. To achieve statistical soundness, at least three experiments were carried out with each sample. The supernatant was analyzed using Perkin–elmer Optima 5300DV Inductively Coupled Plasma Atomic Emission Spectrometer (ICP–AES, U.S.A). Total contents of heavy metals in WWTS or sintered ceramsite were extracted by acid digestion (using HNO 3 /HClO4 /HF) according to USEPA SW3050 and were examined by ICP–AES. Ceramsite were ground in a small agate mortar and XRD patterns of powder ceramsite were recorded on a D/max–γ β X–ray diffractometer with 50mA and 40Kv, Cu Kα radiation (Japan). XRD analyses were conducted by using an XRD pattern database (International Centre for Diffraction Data, ICDD) and the samples were scanned for 2ζ ranging from 10 to 90°. Major components of raw materials and ceramsite containing heavy metals were analyzed by using Philips PW 4400 XR spectrometer (X–ray fluorescence–XRF, Netherlands).

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3. RESULTS AND DISCUSSION 3.1. Part I: Effect of Acidic Oxides (Sio2 And Al2O3) on the Characteristics of Ceramsite 3.1.1. Effect of two acidic oxides on the physical characteristics of ceramsite Figure 1A shows that at a SiO2 content in the raw materials ≤ 30%, the bulk density gradually increases from 510 kg/m3 to 783 kg/m3, while the water adsorption and porosity gradually decrease from 34.9% to 6.9% and from 50.6% to 36.6%, respectively; at a SiO2 content is greater than 30%, the bulk density gradually decreases from 728 kg/m3 to 600 kg/m3, and the water adsorption and porosity gradually increase from 12.4% to 25.2% and from 40.9% to 43.2%, respectively. The maximum particle density (1260kg/m3) is obtained when the content of SiO2 in the raw materials is 27.2%. During the sintering process, the Si4+ is solidified in tectosilicate with network tetrahedron of SiO44- (Si–O–Si). The raw materials with proper contents of SiO2 can enhance the formation of liquid phase at 1000 ℃, which encloses the solid particles and packs the pores in the solid particles. The pores of ceramsite decrease and the binding forces between the solid particles is improved through the reaction described above that also called capillary action (the interaction between contacting surfaces of a liquid and a solid that distorts the liquid surface from a planar shape). Therefore, the variation of SiO2 contents can either reduce or enhance the liquid phase in the ceramsite bodies and make the ceramsite bodies denser or more expanded at 1000 ℃. It is concluded from Figure 1A that 22.5%-45.0% can be considered as the optimal range of SiO2 contents for production of ceramsite. During the sintering process, Al2O3 can be considered as a skeleton material in the sintered ceramsite. Al2O3 can react with other components to form silicate mineral groups (such as anorthite-CaO·Al2O3·2SiO2) with relative lower eutectic points around 1000 ℃ and the reaction effectively lower the sintering point of the materials and enhance the formation of liquid phase [39, 40]. Hence, proper quantities of Al2O3 are beneficial to the liquid-phasesintering at 1000 ℃and improve the characteristics of ceramsite. As shown in Figure 1B, the maximum bulk density (723.9kg/m3) and minimum porosity (42.6%) are obtained when the content of Al2O3 is 15.8%; the maximum particle density (1310kg/m3) and minimum water adsorption (13.6%) are obtained when the content of Al2O3 is 18%. Higher water absorption and lower density of ceramsite (5%≤Al2O3<14% and 26%
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Figure 1. Effect of SiO2 and Al2O3 contents on the physical characteristics of ceramsite (■ bulk density,  particle density, ▲ water adsorption, × porosity).

3.1.2. Thermal properties (DT-TG) analyses The rates of weight loss of the raw materials subjected to sintering were examined at different burning temperatures with the use of a ZRY-2P simultaneous DT-TG analyzer. These tests were conducted in dry air atmosphere to observe the weight loss undergone by the ceramsite samples during sintering. The peaks-valleys on the differential thermal (DT) curves presented the endothermic/exothermic changes during the heating process. TG and DT plots of the samples are closely interrelated. Figure 2A demonstrates that there is little difference in DT-TG curves for mixtures with SiO2 contents of 22.5%, 30%, and 45%. The temperature ranges for significant weight loss for mixtures with SiO2 contents of 22.5%, 30%, and 45% are 235.5-521.4 ℃, 233-520 ℃, and 233.8-522.1 ℃, respectively. The high values of weight loss in these temperature ranges indicate the release of the structural water and a significant amount of organic materials. The maximum slope of weight loss for mixtures with SiO2 contents of 22.5%, 30%, and 45% are demonstrated at 285.5 ℃, 290.6 ℃, and 297.2 ℃. For the 3 mixtures, there are three reaction

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stages in 320-355 ℃, 475-600 ℃, and above 900 ℃ as distinguished by the three distinct endothermic changes over the temperature range. The endothermic reactions observed for the 3 samples above 900 ℃ have been correlated with the formation of a lattice that means a transition from amorphous to a crystalline phase, which in this case implies the crystallization of silicate minerals. 8

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Figure 2. DT and TG analyses for (A) ceramsite with SiO2 contents of 22.5%, 30%, and 45%, and (B) ceramsite with Al2O3 contents of 14%, 18%, and 26%.

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Figure 3. Effect of SiO2 and Al2O3 content on XRD patterns of ceramsite. Band labeling: A, albiteanorthite; E, enstatite; K, kyanite; Q, quartz; S, sillimanite.

As shown in Figure 2B, there are two exothermic peaks for the mixtures with Al2O3 contents of 14%: the peak at 332.6 ℃is due to the removal of absorbed water and the burning of carbon on the surface of ceramsite, and that at 447.7 ℃ corresponds to removal of structural water and carbon in ceramsite bodies. The trend of DT curve for the mixtures with Al2O3 contents of 18% is similar to that of the mixtures with Al2O3 contents of 26%. The continuous endothermic changes for the 3 mixtures begin to occur at 450 ℃, which indicate the endothermic reactions are caused by the gradual transformation of crystalline phases (silicate mineral groups) with lower weight loss as shown in TG curves. These results suggest that the exothermic reaction below 450 ℃ may give ceramsite products with reduced crystallinity and the endothermic reactions above 450 ℃ are slowly conducted that may give the crystals with enhanced stability and strength. However, the composition of raw materials in the optimal ranges of SiO2 and Al2O3 does not dramatically vary, resulting in a predictable characteristic of thermokinetics. The thermokinetics in the actual sintering process can not be properly represented in the DT–TG test, because the shape of ceramsite is spherical and the release of substances may firstly occur on ceramsite surface and then occur in ceramsite interior. Because the combustible organic matters and some of the inorganic matters have to completely release, the DT–TG test still reveals the intrinsic thermokinetics for ceramsite sintering. Results of DT–TG test reinforce the basis of sintering temperature profile adopted in this study (as demonstrated in section 2.2).

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3.1.3. Crystalline phases (XRD) analyses In this study, XRD analyses were applied to obtain mineral compositions of powder ceramsite specimens by using an XRD pattern database (International Centre for Diffraction Data, ICDD). The samples were scanned for 2ζ ranging from 10 to 90°. Using X-ray diffraction, it can be seen from Figure 3 that the changes in the degree of crystallinity of the ceramsite, the formation of new phases, and transformation of some phases. Kyanite (Al2SiO5), quartz (SiO2), and Na-Ca feldspars [albite-Na(AlSi3O8), anorthiteCa(Al2Si2O8)] are the main crystalline phases of the ceramsite with SiO2 contents of 22.5% and a small amount of sillimanite [(Al1.98Fe0.2) SiO5] and enstatite [Mg2(Si2O6)] are also noted in the patterns. The formation of Na-Ca feldspars, sillimanite, and enstatite in ceramsite is caused by the presence of flux such as Fe2O3, CaO, and MgO in raw materials and the solidliquid-phase reactions during the sintering process. Kyanite, quartz, and Na-Ca feldspars are the main crystalline phases of the ceramsite with SiO2 contents of 30%. The major mineralogical compositions found in ceramsite with SiO2 contents of 45% are quartz, kyanite, and Na-Ca feldspars. It should be noted that the broad intensity is attributed to the amorphous silica (between 15° and 30°) and other amorphous phases [39]. The degree of crystallinity of the ceramsite increases as the SiO2 contents increase from 22.5% to 45%. Ceramsite with Al2O3 contents of 14% and 18% consists of both amorphous and crystalline material with a little amorphous material mostly due to silica and the crystalline peaks attributable to kyanite and quartz with small amounts of Na-Ca feldspars and enstatite. Quartz, kyanite, and Na-Ca feldspars with small amounts of sillimanite and enstatite are the identified crystalline phases of the ceramsite with Al2O3 contents of 26%. The main crystalline phases of the 3 ceramsite are similar to each other. During the sintering process, the Al3+ ion may substitute for the Si4+ in a network tetrahedron, contributing to the stability of the network. Therefore, Al2O3 can enter the silica network as AlO44- tetrahedra to replace some of the SiO44- groups; however, since the necessary +4 for the tetrahedra is replaced by the +3 valence of Al, alkali cations must supply the necessary other electrons to produce electrical neutrality (41). Therefore, the presence of Al2O3 in ceramsite results in the formation of numerous crystalline phases (multi-peak pattern) from which aluminosilicates with iron, calcium, magnesium, and sodium have a high probability to exist. The results suggest that Al2O3 plays a significant role in the formation process of crystals, but the major crystalline phases do not dramatically change under the variation of Al2O3 contents.

3.1.4. Morphological structures (SEM) analyses To understand more about the surface morphology and crystalline phases of ceramsite bodies, the samples were gilt with Au and SEM analyses were conducted. The microstructures of the samples were observed using a SEM (S-570 scanning electron microscope) and the crystalline structures are shown in Figure 4. The SEM observations show that the porous structures become more compact due to the increasing SiO2 contents. The observations clearly show the particulate nature of the crystals in the ceramsite with SiO2 contents of 22.5%. Sintered crystals bonding is evident by cohesive necks growing at the particle contact points (Figure 4A1). As seen in ceramsite with SiO2 contents of 30%, there are abundant small pores with thin boundary and big crack with no boundary present on the surface, which are caused by release of gases (Figure 4A2). The

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water absorption of ceramsite with SiO2 contents of 30% remains relatively low despite the more pores, because the gas produced can not create sufficient voids in the ceramsite bodies. Ceramsite with SiO2 contents of 45% clearly shows the advance in densification and neck growth between the particles. However, in Figure 4A3 the ceramsite samples show clear neck growth between the particles but that the particle size is much greater than that of the other two ceramsite, suggesting the occurrence of melted neck growth. The slight expansion of the ceramsite with SiO2 contents of 45% that occurs when sintered at 1000 ℃ is clearly associated with the formation of a significant volume of approximately slit-shaped pores. It can be seen from Figure 4B1 that the surface of the ceramsite with Al2O3 contents of 14% is rough and with a few pores. The micrograph of ceramsite with Al2O3 contents of 18% clearly shows that some pores (6.0μm<pore size<10.0μm) are irregularly distributed in the microstructure and the pores become bigger by comparing with the others (Figure 4B2). Melting phenomena are also observed on the crystalline surface of ceramsite. The release of gases and melting of raw materials allow the big pores to form and the water absorption of ceramsite with Al2O3 contents of 18% remains relatively low despite the bigger pores, because the impervious skin layer of the pellets restricts water ingress. The microstructure of ceramsite with Al2O3 contents of 18% is also related to the lower viscosity of the liquid phase produced at the temperature of 1000 ℃and the consequent improvement on the densification during the natural cooling process. The irregular porous structures with a lot of small pores and agglomerated crystallizations are observed and the microstructure of ceramsite with Al2O3 contents of 26% confirms the formation of several indistinct crystalline phases, as shown in Figure 4B3. The above results indicate that more densified and lower porous ceramsite can be obtained as Al2O3 contents≥18%.

Figure 4. Scanning photomicrographs for ceramsite with SiO2 contents of (A1) 22.5%, (A2) 30%, and (A3) 45%, and ceramsite with Al2O3 contents of (B1) 14%, (B2) 18%, and (B3) 26%.

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Figure 5. Compressive strength analyses for (A) ceramsite with SiO2 contents of 22.5%, 30%, and 45%, and (B) ceramsite with Al2O3 contents of 14%, 18%, and 26%.

3.1.5. Compressive strength analyses The compressive strength is the maximum compressing force a sinter can withstand before it breaks [32, 33, 42]. The sintered ceramsite with diameter of 6-8 mm was placed vertically on the platform of the press (an INSTRON 5569 automatic material testing machine made by INSTRON, USA) and was pressed at a crosshead speed of 0.5 mm min-1 until it was crushed. The compressive strength (N mm-2) of the sintered ceramsite was the compressing force (N) divided by the pressed area (mm2). The compressive strength results were the average values of three tests for each composition. The Chinese National Bureau of Standards (CNBS) requires a minimum compressive strength value of 7.50 MPa (1 MPa=1 N mm-2) for the application of sintered lightweight ceramsite (bulk density< 900 kg m-3) in civil engineering. In this study, results in Figure 5A show that the maximum compressive strength of ceramsite increases from 13.87 to 14.55MPa (above CNBS value) as the SiO2 contents increase. The densified and melting of ceramsite surface are probably the main reasons for the increase in the compressive strength of the ceramsite with higher SiO2 contents (SiO2

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contents≥30%), which implies that textures of ceramsite play an important role in determining the compressive strength. Park and Heo have also been indicated that as the content of SiO2 increase, the bonding strength of metal ions increases, which accordingly can enhance the compressive strength [43, 44]. Another reason is probably the relative increase of the weight ratios of minor components (Fe2O3, CaO, MgO, etc) in the samples because these components are considered to work to lower the strength [44]. But, it should be noted that the effect of SiO2 contents on the compressive strength is not significant. As for different Al2O3 contents in the range of 14%-26%, the compressive strength of sintered ceramsite is well above the value proposed by CNBS (Figure 5B). The compressive strength of ceramsite with Al2O3 contents of 14%, 18%, and 26% are 13.63MPa, 16.25MPa, and 16.15MPa, respectively. Interestingly, their compressive strength seems to be independent of the ceramsite structures. As Al2O3 contents increases from 14% to 18%, higher particle density and more densified bodies are obtained, which inevitably increases the compressive strength of ceramsite. The increasing of Al2O3 contents from 18% to 26% increases the pore rates and water absorption, which may theoretically decrease the compressive strength of ceramsite, but, it is interesting that the engineering properties of the ceramsite thus manufactured are not undermined. The reason may be that during the sintering process, the proper quantities of Al3+ may substitute for the Si4+ in a network tetrahedron [41] and other unreacted Al2O3 present as oxide acts as an accelerator using the readily available alkali cations to increase pozzolanic reactions, contributing to the stability of the network. It is concluded from the above analyses that the strength of sintered ceramsite decreases as Al2O3 contents are below 18%.

3.2. Part II: Effect of Basic Oxides (Fe2O3 and Cao) on the Characteristics of Ceramsite 3.2.1. Effect of two basic oxides on the physical characteristics of ceramsite As shown in Figure 6A, bulk density gradually increases from 497 kg/m3 to 780 kg/m3 as the contents of Fe2O3 increases from 3% to 16%; the maximum particle density (1815.8kg/m3), the minimum water adsorption (8.4%), and the maximum porosity (58.6%) are obtained when the content of Fe2O3 is 10%. The liquid phases can be apparently increased by increasing the contents of Fe2O3 in the raw materials, which enhance the formation of FeO with higher viscosity and lower melting point ( 3Fe2O3 + CO >820℃ 2Fe3O4 + CO2; Fe3O4 + CO >820℃ 3FeO + CO2 ) [45]. But, it should be noted that a sinter with higher viscosity is not good for the release of gases and the formation of pores. 5%-8% is therefore selected as the optimal content range of Fe2O3 for production of ceramsite. Ceramsite has higher bulk density and particle density as the CaO contents are in range of 2.75%-7%, indicating the presence of less intraparticle voids in ceramsite bodies, hence obtaining a lower water adsorption and porosity (as shown in Figure 6B). Higher water adsorption and porosity of ceramsite with lower density may affect the compressive strength and chemical stability of the lightweight ceramsite [12]. 2.75%-7% is selected as the optimal content range of CaO for production of ceramsite.

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Figure 6. Effect of oxides contents on the physical characteristics of ceramsite (A-Fe2O3 and B-CaO). (■ bulk density, □ particle density, ▲ water adsorption, × porosity).

The following tests (thermal analysis, XRD, morphological structures analyses, and compressive strength measurements) were conducted with ceramsite within each optimal contents range of the tested oxides (6%≤Fe2O3≤8% and 2.75%≤CaO≤7%).

3.2.2. Thermal properties (DT-TG) analyses Two exothermic peaks of the mixtures with Fe2O3 contents of 5% are detected at 337.5 ℃ and 473.2 ℃. Two exothermic peaks of the mixtures with Fe2O3 contents of 6% and one exothermic peak of the mixtures with Fe2O3 contents of 8% are detected at 336.7 ℃, 462.5 ℃, and 336.7 ℃, respectively. The intensity of exothermic peaks between 400-500 ℃ decreases as Fe2O3 contents increases and the peak disappears when Fe2O3 contents reach 8%. For each test, the detected exothermic peaks with significant weight loss below 500 ℃ are caused by the release of structural water and mixed gases (CO2, CO, SO2, etc.) (Figure 7A). Endothermic changes are observed and little substances are vaporized from the 3 mixtures above 900 ℃, which indicate the endothermic reaction is caused by the transformation of crystalline phases. Increase of Fe2O3 contents in the mixtures will profoundly influence the thermal properties and lead to more gas produced under the reaction of Fe2O3 at temperatures >820 ℃[45].

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Figure 7. Effect of oxide contents on the thermal properties of ceramsite (A-Fe2O3 and B-CaO).

As shown in Figure 7B, it is interesting that there are two endothermic valleys at 280.6 ℃ and 279 ℃ for the mixtures with CaO contents of 5% and 7%, respectively. These endothermic valleys may partly attribute to the removal of water from hydrated products, which is likely to include most of calcium silicate hydrate (C–S–H) formed by the reaction of CaO, water, and silicate. From the results, it is clear that hydrated products formed at higher CaO contents are one of the first decomposed phases during the heating. There is little difference in DT-TG curves for the 3 mixtures above 300 ℃. Weak weight loss for 3 mixtures are detected above 500 ℃ in TG analyses, which indicate that there is little volatilization of substances and the endothermic changes are caused by the oxidation of mostly inorganic substances and the transformation of crystalline phases [35].

3.2.3. Crystalline phases (XRD) analyses The main crystalline phases of the 3 ceramsite are similar to each other and Quartz (SiO2), kyanite (Al2SiO5), and Na-Ca feldspars [albite- Na(AlSi3O8), anorthite- Ca(Al2Si2O8)] are the main crystalline phases of the 3 ceramsite (Figure 8). The formation of these

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crystalline phases is thought to correspond to the endothermic changes above 900 ℃ as detected in the DT curves (Figure 7A). The enstatite [Mg2(Si2O6)], sillimanite [(Al1.98Fe0.2)SiO5], and ferrosilite magnesian [(Fe,Mg)SiO3] have a high probability to exist in the ceramsite with Fe2O3 contents of 8%, which can be attributed to the melting and diffusing of abundant Fe2O3 at 1000 ℃. Quartz, kyanite, and Na-Ca feldspars are the main crystalline phases of the ceramsite with CaO content of 2.75% (Figure 8). Na-Ca feldspars has already been formed in ceramsite with CaO content = 2.75%, as one of the minor crystalline phase, co-existing with quartz and kyanite. The most pronounced crystalline phases of the ceramsite with CaO contents of 5% and 7% are Na-Ca feldspars, kyanite, and Quartz with a small amounts of enstatite [Mg2(Si2O6)] and sillimanite [(Al1.98Fe0.2)SiO5]. As the CaO contents increase from 2.75% to 7%, the intensity of XRD peaks of Na-Ca feldspars increases while the peaks of quartz and kyanite crystals tend to decrease and the amorphous phases increase in the ceramsite. As the total amount of quartz decreases, Na-Ca feldspars derived from kyanite dissolution either remain in glassy phase or are used to nucleate other crystals in ceramsite [35]. Better crystallization produces better strength and chemical durability [45], which implies that the compressive strength of ceramsite may decrease as CaO contents increase from 5% to 7%.

3.2.4. Morphological structures (SEM) analyses It is observed that the ceramsite with Fe2O3 contents of 5% presents rough and densified surface and has some irregularly distributed approximately slit-shaped pores. It can be explained that the pores form as the residual glassy phase viscosity falls to a level when gasforming inorganic decomposition reactions can produce the observed pores [44]. Ceramsite with Fe2O3 contents of 6% and 8% present rough surfaces with a few small pores. These are believed to be a result of the softening of the glassy phase present in the ceramsite, along with a simultaneous incomplete release of gas at 1000 ℃. Comparing the images in Figure 9 (A1A3) reveals that the higher the Fe2O3 contents, the smaller and fewer the pores. The SEM observations for the 3 ceramsite are in general agreement with the trends shown by the physical data (Figure 6A). The above results indicate that ceramsite has better crystallization and melting of the bodies as 6%≤Fe2O3≤8%. As shown in Figure 9B1, ceramsite with CaO contents of 2.75% presents rough and densified surface and has a few irregularly distributed pores. Crystals with many-sided morphology and a few small pores are observed on the surface of ceramsite with CaO contents of 5%. For ceramsite with CaO contents of 2.75% and 5%, melting phenomena are observed on both of the crystalline surfaces. SEM image of ceramsite with CaO contents of 7% suggests that the majority of the coarse and fine minerals are located in close contact with the silicates matrix. The relative higher alkaline earth oxide content (CaO), present in the ceramsite will act as a fluxing agent during the sintering process [46, 47]. Besides better porosity, ceramsite surfaces are glossier when densification is more completed at lower CaO contents. This might partly explain the differences in compressive strength for the 3 samples because of ceramsite with denser surface generally have better strength. The differences in the microstructures (Figure 9B1, 9B2, and 9B3) are also suggested to be the reasons for the differences in their water absorption, because the size and quantity of pores are the chief determinants of water absorption.

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2theta Figure 8. Effect of oxide contents on the XRD patterns of ceramsite. Band labeling: A, albite-anorthite; E, enstatite; F, ferrosilite magnesian [(Fe,Mg)SiO3]; K, kyanite; Q, quartz; S, sillimanite [(Al1.98Fe0.2)SiO5].

Figure 9. Effect of oxide contents on the morphological structures of ceramsite (A-Fe2O3 and B-CaO).

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Figure 10. Effect of oxides contents on the compressive strength of ceramsite (A-Fe2O3 and B-CaO).

3.2.5. Compressive strength analyses The compressive strength of ceramsite with Fe2O3 contents of 5%, 6%, and 8% are 14.97MPa, 15.14MPa, and 15.67MPa, respectively (Figure 10A). Ceramsite with Fe2O3 contents of 5% has irregular-shaped elongated pores that decrease the compressive strength. Because of its porous structure, which is caused the foaming reactions and the tiny pores produced during sintering, the resultant specimens have worse compressive strength property. In contrast, ceramsite with a few pores formed at higher Fe2O3 contents lead to relatively higher compressive strength. The difference in crystalline phases may also explain the higher gain in compressive strength for ceramsite with 8% Fe2O3 as compared to ceramsite with 5% and 6% Fe2O3. The Fe3+ may act as Al3+, replacing the Si4+ parent ions of silicates to be enclosed in the framework of silicates and lower the body strength. It can be seen from Figure 10B that the compressive strength of ceramsite with CaO contents of 2.75%, 5%, and 7% are 15.13MPa, 14.26MPa, and 13.13MPa, respectively. Ca2+ may act as metal ion ionically bonded in the interstices of the silicate network to produce electrical neutrality that was broken due to the substitution of Si4+ by Al3+ [39]. Increasing the content of CaO usually makes the crystalline particles easier to form at a given temperature but increases its chemical reactivity in the silicate network. The compressive strength decreases as the content of CaO increase from 2.75% to 7%, which implies that excessive CaO exceeds the needed ions for producing electrical neutrality and lowers the body strength. Another reason may be that the lower strength is thought to be due to the lower hardness of the resulting crystalline phases [45].

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Figure 11. Effect of sintering temperature on the leaching of heavy metals.

3.3. Part III: Stabilization of Heavy Metals in Ceramsite 3.3.1. Effect of sintering temperature on stabilization of heavy metals To indicate the actual resistance provided by the structures of ceramsite for leaching of heavy metals, the effect of sintering temperature on the solidification of heavy metals was investigated. It can be seen from Figure 11 that the leaching contents of Cd, Cr, Cu, and Pb of the 3 specimens decrease at the 24th hour or on the 30th day as sintering temperature increases. The leaching contents of Cd and Cu of the 3 specimens at the 24th hour or on the 30th day are all near 0μg/g and almost do not change above 975 ℃. The increase in sintering temperature above 1000 ℃ has a slight influence on the leaching contents of Cr and Pb. It should be also noted that Cd and Cu are the elements most sensitive to sintering temperature (≥1000 ℃), and the leaching contents of the 4 heavy metals on the 30th day are higher than those at the 24th hour. It is demonstrated that the heat treatment of ceramsite is advantageous to improve the thermodynamic stability of heavy metals but that, to some extent, thermal treatment at temperature below 1000 ℃reduces the binding capacity for heavy metals. Above results indicate that after subjecting the sintered ceramsite to leaching conditions that sintering at≥1000 ℃exhibits good binding ability for Cd, Cr, Cu, and Pb in ceramsite and can be considered as a proper means for solidifying/stabilizing heavy metals. At temperatures below 1000 ℃, densification and crystals growth are the main processes occurred in ceramsite and heavy metals can not be completely solidified in ceramsite with the expanded bodies and

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the semideveloped crystalline phases. This implies that only weak physical bonds are formed between these heavy metals and the compounds in the ceramsite, and thus heavy metals are easy to be leached, which may pose a detrimental impact on the environment. More densified and lower porous ceramsite can be obtained at temperatures ≥1000 ℃and the high temperatures are beneficial to the formation of liquid phases and improve the strength characteristics of ceramsite [48], which inevitably increase the solidification efficiencies of heavy metals in ceramsite. Furthermore, at higher temperatures, the heavy metals in the raw materials can be entrapped inside the crystalline structures through the chemical and physical transformation. It is generally recognized that the leaching of heavy metals obtained by such leaching procedures is always operationally affected by many experimental factors especially the mineral compositions of the samples [48]. In this study, XRD analyses were applied to obtain mineral compositions of powder ceramsite specimens by using an XRD pattern database (International Centre for Diffraction Data, ICDD). Samples sintered at 1000 ℃ with different added contents of heavy metals (minimum, medium, and maximum) were selected and the samples were scanned for 2ζ ranging from 10 to 90°. The analyses are also helpful to test the forms of heavy metals present in ceramsite in order to judge the actual improvement in metal binding resulting from the thermal treatment. It can be seen from Figure 12A that the main crystalline phases of the 3 kinds of ceramsite with different heavy metals contents are similar. The results indicate that the effect of the contents of heavy metals on the formation of main crystalline phases in ceramsite is minor. Kyanite (Al2SiO5), quartz (SiO2), and Na-Ca feldspars [albite-Na(AlSi3O8), anorthiteCa(Al2Si2O8)] are the main crystalline phases of the 3 ceramsite specimens and a small amount of sillimanite [(Al1.98Fe0.2) SiO5] and enstatite [Mg2(Si2O6)] are also noted in the patterns. The formation of Na-Ca feldspars, sillimanite, and enstatite in ceramsite is caused by the presence of flux such as Fe2O3, CaO, and MgO in the raw materials and the solid-liquidphase reactions during the sintering process. Better crystallization produces better strength and chemical durability [49], which implies that the solidifying efficiencies of heavy metals in ceramsite may be partly influenced by the formation and types of crystals. It is thus evident that the fraction of metals available for leaching is greatly reduced as a result of the crystallization and chemical fixation within the aluminosilicates or silicates frameworks acting during thermal treatment, as reported by the previous studies [48, 50]. The evidence of chemical immobilization of metals is gained from the mineralogical characterization, in that the heavy metals compounds in the 3 specimens can be identified in ceramsite by XRD analyses (as shown in Figure 12B). Figure 12B shows that the 4 heavy metals in ceramsite sintered at 1000 ℃ are in steady forms and the main compounds are crocoite (PbCrO4), chrome oxide (Cr2O3), cadmium silicate (CdSiO3), and copper oxide (CuO). The results indicate that the interactions between the chromate and insoluble Pb compounds (e.g., lead oxides) result in the formation of crocoite (PbCrO4), revealing the low leachability of Pb. Thus, the solidifying efficiency due to the formation of PbCrO4 via consumption of the CrO42- present in the raw materials is expected to be significant and the sintering process enhances Pb2+ to form insoluble PbCrO4 with CrO42- [51]. It is also observed that some Cr6+ is deoxidized to Cr3+ reflecting by the identification of Cr2O3 in XRD patterns. The formation of CuO provides direct evidence that Cu can be trapped inside the newly formed silicates or aluminosilicates minerals and it is reasonable to think that Cu is not located in the framework, but is most likely buried inside the newly formed minerals.

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Figure 12. XRD analyses of the main crystalline phases (A) (Band labeling: A, albite-anorthite; E, enstatite; K, kyanite; Q, quartz; S, sillimanite) and the forms of heavy metals (B) (Band labeling: ■, CuO; ▼, PbCrO4; ▲, CdSiO3; ●, Cr2O3) in ceramsite sintered at 1000 ℃.

The forms of Cr, Cu, and Pb (i.e. Cr2O3, CuO, and PbCrO4) particularly suggest that more incorporation of the compounds into the aluminosilicates or silicates matrix is occurred after the heat-induced transformation. For Cr, Cu, and Pb, this transformation of the chemical state of the incorporated metals is clearly supported by the leaching results and XRD studies. Majority of Cd is uniformly distributed in ceramsite bodies to form CdSiO3 indicating that Cd has entered the liquid-solid-phases and stabilized in the crystalline structures of ceramsite during the sintering process. The transformation of these heavy metals to crystalline state is advantageous for the long-term stability of the metals and the crystalline solids are likely to have the improved capacity to bind heavy metals [50]. Moreover, it is possible that the heatinduced transformation of crystallization causes most of the heavy metals ions to transfer from the surface to the interior of the particles. It is thus concluded that the fraction of heavy

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metals available for leaching is greatly reduced and the solidifying efficiencies are strongly enhanced by the crystallization and chemical incorporation within the aluminosilicates or silicates frameworks during the thermal treatment.

3.3.2. Effect of pH on leaching behaviors of heavy metals The pH-dependent and the following H2O2-dependent leaching tests of heavy metals are conducted with ceramsite sintered at 1000 ℃ to investigate the effect of external conditions on the stabilization of heavy metals. It can be seen from Figure 13 that the maximum leaching contents of the 4 heavy metals is achieved at the 24th hour or on the 30th day when pH is 1 and the leaching contents almost do not change as pH>1. It is interesting to note that the cumulative amounts of Cr leached from ceramsite (1000mg kg-1) on the 30th day are the maximum and about fourfold as much as those at the 24th hour. Although the above phenomenon is present, it should be noted that Cr cannot be easily released from ceramsite during the pH-dependent leaching test. Most of Cr6+ may be deoxidized to Cr3+ (Cr2O3) in the reducing environment formed locally during the sintering process and both Cr6+ and Cr3+ can be stabilized in the structures of crystalline networks [41]. The pH-dependent leaching tests indicate that the relationship between heavy metals mobility and pH is complicated and the solubility of heavy metals in ceramsite are dramatically influenced by lower pH such as pH=1, also reported in previous studies [36]. In the leaching tests, the pores in ceramsite are slowly and sufficiently filled by the leachant so that the pores water pH will progressively decrease especially when pH is 1 and then the heavy metals in this region will start to leach. A fraction of the leached metals will diffuse toward the leachant and these initial released metals likely originate from the outer surface pores of ceramsite in contact with the leachant. However, the results from the present work indicate that the residual amounts of heavy metals in the sintered ceramsite appeared to be efficiently immobilized within the silicates or aluminosilicates matrix, as demonstrated by the availability values as expressed in terms of the leaching contents of each heavy metal. This implies that stronger chemical bonds are formed between these heavy metals and the components in ceramsite, making heavy metals difficult to be leached, in other words, the toxic metals present in the ceramsite pose no harmful impact on the environment.

3.3.3. Effect of oxidative condition on leaching behaviors of heavy metals It can be seen from Figure 14 that the leaching contents of Cd, Cr, Cu, and Pb at the 24th hour or on the 30th day increase as the H2O2 concentration increases. This phenomenon indicates that heavy metals can be significantly solidified in ceramsite for a long period of time due to the heavy metals in the pores surface present in stable forms and will remain steadily even in the oxidative environment. Furthermore, the leached space of the pores hinders the initiation of more contact oxidation, which decreases the available space for water ingress and improves the solidification efficiencies of Cd, Cu, and Pb in ceramsite. The leaching contents of heavy metals slowly increase at H2O2 concentration≥1.5mol L-1, which can be attributed to the reason that the solubility of heavy metals and oxidative ability of H2O2 have already reached an equilibrium state in the leachant.

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The reasons for the effective solidification of heavy metals in oxidative condition can be attributed to many phenomena and are presented as follows: During the sintering process, the Al3+ ion may substitute for the Si4+ in a network tetrahedron, contributing to the stability of the network. Therefore, Al2O3 can enter the silica network as AlO44- tetrahedra to replace some of the SiO44- groups; however, since the necessary +4 for the tetrahedra is replaced by the +3 valence of Al, cations must supply the necessary other electrons to produce electrical neutrality [41]. These complex chemical reactions allow for the incorporation of a large number of cations including Cd, Cr, Cu, Pb, and others into the structure [48, 52]. Therefore, these results indicate that, even subjecting the ceramsite to rigorous leaching conditions (H2O2 concentration=5mol L-1), ceramsite structures systems still exhibit good binding ability for heavy metals and cannot be easily leached from ceramsite to the leachant, which can be considered chemically durable. It can be concluded from the results that most of the heavy metals are incorporated inside the crystalline structures of ceramsite and these new recycled materials are environmentally sustainable. It seems that higher contents of Al2O3 may enhance the (Cu and Pb) substitution of parent ions (Al3+ or Ca2+) in ceramsite and may restrain the replacement of parent ions by Cd and Cr. This phenomenon indicates that heavy metals can be significantly solidified in ceramsite for a long period of time because the heavy metals with stable forms in the porous surface can not be easily washed out. It should be noted that Cu leached from ceramsite with lower Cu contents on the 30th day is higher than those with higher Cu contents, which may be attributed to the special reaction between Cu and the components in ceramsite. Because of the enhancement of diffusion for the ions of Cu in the sintering processes, cation-exchange reactions between the ions of Cu and the crystals in ceramsite are easily to occur and, consequently, cation exchange is partly responsible for the immobilizations of such heavy metal ions (Cu). The higher Cu contents in ceramsite may enhance their substitution of parent ions (Al3+ and Ca2+) and may therefore enable the leaching content of Cu to decrease [23].

4. CONCLUSION It can be concluded from the results in this chapter that DWTS can be used as a substitute for clay to be mixed with WWTS and water glass for producing ceramsite with optimal content of SiO2 and Al2O3 ranging from 14–26% and 22.5–45%, respectively. Thermal properties are profoundly influenced by the variation of Fe2O3 contents in the mixtures. Ceramsite with higher Fe2O3 contents have more complex crystalline phases and fewer pores, which accordingly contribute to relatively higher compressive strength. As the content of CaO increases, pores, Na-Ca feldspars, and amorphous phases increase while compressive strength decreases, which implies that excessive Ca2+ exceeds the needed ions for producing electrical neutrality of silicate networks. This study also demonstrates the feasibility of transforming heavy metals into stable forms in ceramsite and the successful reduction of heavy metals mobility after the heat treatment. When the sintering processes are conducted with the ceramsite containing heavy metals, a decrease in the solubility of the heavy metals can be achieved due to the changes in their structural location and chemical forms. Sintering temperature and the changes in composition of leachant solution (acidic, neutral, alkaline and oxidative) have significant influence on the leaching and solidification behaviors of the heavy

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metals in ceramsite. Heavy metals stabilized in ceramsite are in steady forms and cannot be easily released into the environment again to cause secondary pollution. The results in this chapter present a new step towards the safe utilization of sludge-ceramsite. The findings of this study can revolutionize the handling of such kinds of sludge in the future for their reuse as low-cost materials, rather than their becoming waste requiring costly disposal.

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[31] Muruganandham, M., Chen, S. H. & Wu, J. J. (2007). Evaluation of water treatment sludge as a catalyst for aqueous ozone decomposition. Catal. Commun., 8, 1609-1614. [32] Huang, C. P., Pan, J. R. S. & Liu, Y. R. (2005). Mixing water treatment residual with excavation waste soil in brick and artificial aggregate making. J. Environ. Eng., 131, 272-277. [33] Tay, J. H., Hong, S. Y. & Show, K. Y. (2000). Resue of industrial sludge as palletized aggregate for concrete. J. Environ. Eng., 126, 279-287. [34] Xu, G. R., Zou, J. L. & Dai, Y. (2006). Utilization of dried sludge for making ceramsite. Water Sci. Technol., 54, 69-79. [35] Xu, G. R., Zou, J. L. & Li, G. B. (2008). Effect of sintering temperature on the characteristics of sludge ceramsite. J. Hazard. Mater, 150, 394-400. [36] Xu, G. R., Zou, J. L. & Li, G. B. (2008). Stabilization of heavy metals in ceramsite made with sewage sludge, J. Hazard. Mater, 152, 56-61. [37] Xu, G. R., Zou, J. L. & Li, G. B. (2008). Solidification and leaching behaviors of Cr6+ in sludge ceramsite, J. Hazard. Mater, 153, 1031-1035. [38] Xu, G. R., Zou, J. L. & Li, G. B. (2008). Ceramsite made with water and wastewater sludge and its characteristics affected by SiO2 and Al2O3. Environ. Sci. Technol., 42, 7417-7423. [39] Acchar, W., Vieira, F. A. & Hotza D. (2006). Effect of marble and granite sludge in clay materials. Mater. Sci. Eng., 419, 306-309. [40] Shih, K., White, T. & Leckie, J. O. (2006). Spinel formation for stabilizing simulated nickel-laden sludge with aluminum-rich ceramic precursors. Environ. Sci. Technol., 40, 5077-5083. [41] Li, C. T., Lee, W. J., Huang, K. L., Fu, S. F. & Lai, Y. C. (2007). Vitrification of chromium electroplating sludge. Environ. Sci. Technol., 41, 2950-2956. [42] Yilmaz, O., Ünlü, K. & Cokca, E. (2003). Solidification/ stabilization of hazardous wastes containing metals and organic contaminants. J. Environ. Eng., 129, 366-376. [43] Park, Y. J. & Heo, J. (2002). Vitrification of fly ash from municipal solid waste incinerator. J. Hazard. Mater, B91, 83-93. [44] Toya, T., Kameshima, Y., Yasumori, A. & Okada, K. (2004). Preparation and properties of glass-ceramics from wastes (Kira) of silica sand and kaolin clay refining. J. Eur. Ceram. Soc., 24, 2367-2372. [45] Mansfeldt, T. & Dohrmann, R. (2004). Chemical and mineralogical characterization of blast-furnace sludge from an abandoned landfill, Environ. Sci. Technol., 38, 5977-5984. [46] Ducman, V., Mladenovic, A. & Suput, J. S. (2002). Lightweight aggregate based on waste glass and its alkali-silica reactivity, Cement Concrete Res., 32, 223-226. [47] Barbieri, L., Bonamartini, A. C. & Lancellotti, I. (2000). Alkaline and alkaline-earth silicate glasses and glass-ceramics from municipal and industrial wastes, J. Eur. Ceram. Soc., 20, 2477-2483. [48] Ho, M. D. & Evans. G. J. (2000). Sequential extraction of metal contaminated soils with radiochemical assessment of readsorption effects. Environ. Sci. Technol., 34, 1030-1035. [49] Toya, T., Nakamura, A., Kameshima, Y., Nakajima, A. & Okada, K. (2007). Glassceramics prepared from sludge generated by a water purification plant. Ceram. Inter., 33, 573-577.

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[50] Cheng, T. W. & Chen, Y. S. (2003). On formation of CaO–Al2O3–SiO2 glass–ceramics by vitrification of incinerator fly ash. Chemosphere, 51, 817-824. [51] Fang, Z., Xu, S. k. & Koziński, J. A. (2000). Behavior of metals during combustion of industrial organic wastes in supercritical water. Ind. Eng. Chem. Res., 39, 4536-4542. [52] Martinez, C. E. & McBride, M. B. (2000). Aging of coprecipitated Cu in alumina: Changes in structural location, chemical form, and solubility. Geochim. Cosmochim. Acta, 64, 1729-1736.

In: Fluid Waste Disposal Editor: Kay W. Canton, pp. 111-135

ISBN: 978-1-60741-915-0 © 2010 Nova Science Publishers, Inc.

Chapter 4

MODELLING AND OBSERVATION OF PRODUCED FORMATION WATER (PFW) AT SEA

a

D. Cianellia,b, L. Manfrab, E. Zambianchia, C. Maggib and A. M. Cicerob Parthenope University of Naples, Department of Environmental Sciences, Centro Direzionale Isola C4, 80143 - Naples, Italy. b ISPRA – Institute for Environmental Research and Protection, Via di Casalotti 300, 00166 - Rome, Italy.

ABSTRACT Through the last decades, the ever increasing energetic demands have been accomplished by exploiting new natural reservoirs, including offshore oil and gas deposits located in marine coastal areas. During the extraction and production phases, large amounts of water are brought to the surface along with the hydrocarbons. These waters include the ‗formation water‘, that lies underneath the hydrocarbon layer, and ‗additional water‘ usually injected into the reservoirs to help force the oil to the surface. Both formation and injected waters, named ―produced formation waters‖ (PFWs), are separated from the hydrocarbons onboard offshore platforms and then disposed into the marine environment through ocean diffusers. PFWs contain several contaminants and represent one of the main sources of marine environment pollution associated with oil and gas production. This makes the study of PFW fate of paramount importance for a proper management of environmental resources as well as for planning and optimizing the discharge and monitoring procedures. In the first part of this chapter we provide a detailed description of the chemical characteristics of PFWs and their potential toxic effects and review the mixing processes governing their dispersion in the marine environment. In the second part of the work we briefly review past efforts in observing and modelling PFW spreading in the ocean. Finally, we propose a multidisciplinary approach, integrating in situ observations and numerical modelling, to assess dispersion of PFWs in space and time. As a case study we will refer to the results of a previous study conducted in the Northern Adriatic Sea, a subbasin of the Mediterranean Sea, where a number of offshore natural gas (CH4) extraction platforms are currently active.

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INTRODUCTION The worldwide exploitation of oil and gas reserves located in marine coastal areas has drastically increased during the last decades due to the gradual depletion of the on-land deposits. A by-product resulting from both oil and gas extraction is the Produced Formation Water (PFW), which to date represents the largest volume waste from the production phase of the offshore oil and gas facilities (e.g. Ray and Engelhardt, 1992; Berry, 2005). The disposal of the PFWs is carried out through the reinjection into the reservoir, the discharge into the ocean or the transport onshore. Here we focus on the disposal of the PFWs into the marine environment by means of surfacing or submerged outfalls. The complex chemical composition of the PFWs (ranging from heavy metals to soluble hydrocarbons) as well as its physical state (solution, particulate-matter, suspensions, etc.) determines the potential impact of the PFWs on the receiving environment (OGP, 2005). In the last decade the release of PFW into the sea has received increasing attention due to both the potential long term effects (and associated ecological risks) of some of its chemical compounds and the high water volumes discharged (Rye et al.,1996). In order to determine the actual impact of the PFWs on the marine environment, the chemical and physical processes driving their distribution into the ambient fluid have to be investigated. In particular, an exhaustive assessment of the effect of PFWs on the environment at discharge locations may be derived from the site-specific chemical characteristics of the effluent and environmental factors (meteo-oceanographic conditions). After the release into the sea, the PFW undergoes several different processes such as the dilution into the ambient fluid (which may occur more or less rapidly depending on local oceanographic conditions), the evaporation and biodegradation (that change the concentration and composition of inorganic constituents) (Strømgren et al., 1995 and references therein), the volatilization towards the atmosphere and the settling at the bottom. In particular, a crucial process determining the dilution of PFW and the reduction of its concentration in the sea water is the mixing of the effluent with the ambient fluid (e.g. Baumgartner et al., 1994). Such a dispersion process occurs in two phases: a rapid initial mixing phase (near-field) taking place immediately after the release within the first tens of meters from the discharge point, and the subsequent passive dispersal phase (or far-field) that evolves at larger distance over time scales of hours or days. Nowadays several field measurement and dispersion modelling studies investigating the fate of the PFW and predicting the near and far field dilution rates with high accuracy are available (e.g. Neff, 2002). The approach integrating field and laboratory observations with numerical modelling presently represents the most suitable tool to assess the effects of PFW discharges and the potential risks for the local marine ecosystems. Such studies can be considered as a baseline to improve monitoring programs and to optimize different sampling, monitoring and assessment techniques as well as the industry discharge practices (Durell et al., 2006; Cianelli et al., 2008). In this paper we firstly describe origin, chemical characteristics, potential toxic effects and methods of disposal of the PFWs. In the first part we also provide a description of the dispersion processes that a marine discharge may experiment and the possible effect of environmental factors (water column stratification and ambient current) on the effluent.

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Thereafter we present a brief review of field observations and numerical modelling studies on the fate of PFWs in the marine environment. In the last section we summarise the results of a recent work conducted in the Northern Adriatic Sea, a sub-basin of the Mediterranean Sea, where a number of offshore natural gas (CH4) extraction platforms operate (Cianelli et al., 2008). This work represents a case study applying a multidisciplinary approach to investigate the initial mixing of PFW discharged from three offshore natural gas platforms. The research compares field and laboratory observations with the near field dilution predicted by a numerical model. The results show the temporal and spatial distribution of the PFW and provide useful insights to improve the PFW discharge monitoring in the Adriatic Sea.

CHEMICAL CHARACTERISTICS, POTENTIAL EFFECTS AND DISPOSAL METHODS OF THE PFWS The natural water that lies underneath the hydrocarbons in oil/gas reservoirs is called ―formation water‖. During extraction operations water is injected into the reservoir to help force the oil/gas to the surface. The water produced from the well is called ―Produced Formation Water‖ (PFW) (Environment Australia, 2001). Its composition is complex: studies have been focussed mainly on PFWs originating in the North Sea (see Ray and Engelhardt, 1992 and Reed and Johnsen, 1996; for Adriatic PFW composition see Maggi et al., 2006; Manfra et al., 2007; Mariani et al., 2004). The PFW composition depends on the geological characteristics of the reservoir (e.g., Utvik, 1999). The composition also changes according to production: water from gas production fields has generally a higher content of low molecular-weight aromatic hydrocarbons than water produced from oil fields. For this reason, PFW originating from gas platforms might be more toxic compared to water produced in oil installations (Jacobs et al., 1992). However, gas platforms tend to produce small water volumes compared to oil platforms and this may be an important element to take into account when it comes to mitigating the impact of PFW discharge (OGP, 2005). These aspects may in part explain the differences between PFW originating in different marine areas, for example the North Sea (mainly oil fields) and the Adriatic Sea (principally gas platforms). PFWs contain organic compounds such as dispersed oil (in droplet form), dissolved oil, organic acids and phenols. In particular, volatile aromatic compounds as Benzene, Toluene, Ethylbenzene, Xylenes (BTEX) and low molecular weight aromatic hydrocarbons as Naphthalene, Phenanthrene and Dibenzothiophene and their alkyl homologues (NDP) are mainly in dissolved form (OGP, 2002), while high molecular weight Polycyclic Aromatic Hydrocarbons (PAHs) are present as dispersed oil (Patin, 1999). They may be removed from seawater by means of volatilisation, adsorption to particles, sedimentation, biodegradation and photolysis. Exposure to marine organisms is low and may cause narcosis, alterations of permeability of cell membranes and developmental defects (OGP, 2005). PAHs are relatively insoluble and their potential for bioaccumulation increases with increasing molecular weight. They may be toxic in different way as narcosis, phototoxicity, biochemical activation, etc. (OGP, 2005). Phenols are soluble in marine waters, highly volatile and degradable; they appear not particularly toxic to aquatic organisms.

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PFWs also include inorganic compounds as chlorinates, carbonates, sulphurs, sodium, potassium, calcium, ammonium and trace metals (copper, nickel, iron, chromium, manganese, lead, zinc, barium, arsenic, mercury and cadmium). The most common trace elements are barium, iron, manganese, zinc and lead. Some authors suggested that corrosion of galvanised equipment could be a source of zinc and lead in PFW (Neff, 1992; OGP, 2005). PFWs show metal concentrations higher than seawaters, in particular for those produced in gas reservoirs, but often is not possible to distinguish the PFW contributes in the marine environment due to dilution and physicochemical reactions (co-precipitation, adsorption, etc.). PFWs may also contain chemical additives as biocides, inhibitors of corrosion and scale, inhibitors of emulsion and reverse emulsion, coagulants/flocculants and antifoaming agents (Brendehaug et al., 1992). Some of the additives are water-soluble, i.e. diethylene glycol used on offshore gas platforms to prevent the hydrate formation during the gas–water separation process and to inhibit corrosion events. Their presence in PFW may have an effect on oil/water partition coefficient, acute and chronic toxicity, bioavailability and biodegradability. PFWs are disposed by rejection, marine discharge or transport onshore. Mechanical oil/water separation devices and chemical treatments are used to separate oil and water before PFW is discharged into the sea. In spite of treatments, PFW still contains contaminants and for this reason its discharge is strictly regulated and controlled. The offshore protocol of Barcelona Convention established a limit concentration of 40 mg/litre for oil content in PFW. The Italian Law establishes a limit of 40 mg/litre for oil content in PFW and imposes the monitoring of possible impacts on marine ecosystems induced by the PFW discharge.

FATE OF PFWS IN THE MARINE ENVIRONMENT Weathering Processes When PFW is discharged into the sea, its chemical compounds undergo a number of dispersion, removal and degradation processes (Burns et al., 1999). In particular, weathering is a combination of physical, chemical and biological processes that influence the fate of PFW discharged into the sea and consequently the distribution and the effects of compounds present in it. The most important sub-processes of weathering are dilution (discussed in the next section), evaporation, chemical reactions, adsorption, sedimentation and biodegradation. Evaporation of PFW-contained contaminants depends on vapour tension of compounds and some environmental conditions (e.g. current velocity, water depth, turbulence, wind velocity). The compounds with lower molecular weight and higher vapour tension tend to evaporate into the air rapidly. Some aromatic hydrocarbons (like, e.g., BTEX) may be dissolved in seawater but easily volatilize into the air. The compounds of PFW are also exposed to different chemical reactions (e.g. oxidation, hydrolysis and precipitation). Some PFW elements (e.g. trace metals) exist at lower oxidation states until not in contact with oxygen; they undergo oxidation when PFW is brought to the surface and then precipitate as hydroxides. As an example consider the transformation of dissolved iron (Fe2+) to Fe3+ and to hydroxides or the oxidation of hydrogen sulphide to elemental sulphur that precipitates (OGP, 2005).

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Table I. Main processes acting on PFWs (Roberts, 1990). Phase Near field Far field Long term

Process Initial mixing Buoyancy and momentum Advection – oceanic currents Diffusion –oceanic turbulence Large scale circulation Upwelling / downwelling Sedimentation

Spatial scale (m) 10 – 1000

Temporal scale 1 – 10 min.

100 - 10000

1 – 20 h

10000 - 100000

1 -100 d

Some substances tend to adsorb on suspended solids both within PFW and in the water column; in this way, they become potentially available to filter feeders and/or transported into the sediments as a result of flocculation and complexation processes; the adsorption depends on factors such as the polarity and solubility of chemical substances, the particle size (preference for fine fraction <60µm) and organic carbon content in particulates and sediments (Marchetti, 2000). Sedimentation yields accumulation of these compounds in benthic organisms and may result in inserting them into the trophic chain. PFW compounds can be biodegraded from the microbial community both in seawater and in the sediment; aerobic processes take place in the water column while in the sediment aerobic or anaerobic degradation can occur. The biodegradation is conditioned by factors such as the chemical structure of compounds, the availability of micro-organisms and the environmental conditions (e.g., oxygen and nutrient concentrations); substances with low molecular weight are readily broken down while multi-ring aromatic hydrocarbons are relatively stable and remain longer in the plume of PFW (OGP, 2005).

Mixing of Effluents Discharged into the Marine Environment After the separation from hydrocarbons, PFWs are discharged directly into the sea from offshore platforms by means of surface or submerged outfalls, usually mounted in a vertical position. The effluent, also indicated as plume, is released at a certain depth and undergoes physical, biological and chemical processes acting over increasing temporal and length scales while moving away from the release point. Among these processes, the most important ones together with their typical scales are reported in Table I. The mixing in the proximity of the diffuser is influenced by buoyancy and momentum fluxes, which are generated by the release of PFWs from the diffuser itself and by their interaction with the current field (self-induced turbulence). The buoyancy is originateds by the density difference between the effluent and the ambient fluid, while the momentum is associated with the velocity at which the effluent is ejected. The interaction with the medium is intense and determines a fast mixing of the two fluids. This process persists until the turbulent kinetic energy generated by these processes is dissipated. The region where this initial mixing occurs is named ―near field‖ or ―initial mixing zone‖ or alternatively ―initial dilution zone‖; the outer limit of this area is generally

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set where the plume meets a boundary layer (e.g., the sea surface, the bottom of the water column or an equilibrium layer where the effluent density equals that of the environment). By the end of the initial mixing, the effluent is stabilised and passively transported by the ocean currents in an area known as the ―far field‖, where the mixing is entirely due to the environmental turbulence and takes place more slowly than in the near field. At distances greater than that of the far field, PFWs may be subject to chemical and biological decay processes determining their long term fate (Roberts, 1990).

The dilution factor Generally, an effluent is made of a transport fluid (e.g., water) carrying undesired substances, the pollutants. In this case, the dilution process is based on the mixing of the transport fluid with the ambient one, typically water mixing with other water. The mixing process by which the ambient fluid incorporates the effluent is known as entrainment. The dilution process can be assumed as a by-product of entrainment. To quantify the dilution process, a parameter is commonly used, the average dilution factor Sa:

Sa =

ve  va ve

(1)

where ve is the effluent flux volume discharged, while va is the ambient fluid flux volume (Baumgartner et al., 1994). In the vicinity of the diffuser port, Sa will be close to 1 as va will be very small. Accurately measuring va is not straightforward; for this reason, very often what is measured is the concentration of pollutant in a selected position at a given distance from the discharge. The mass balance equation thus gives:

cp ve  va   ceve  cava

(2)

where cp is the average concentration of the effluent cross section, ce is the initial concentration in the effluent and ca that in the ambient fluid. Reorganising equation #2, Sa becomes (Glenn, 1997):

Sa 

ce  ca cp  ca

(3)

If the ambient concentration is null (ca = 0), the equation #.3 simplifies as (Baumgartner et al., 1994):

Sa  ce cp

(4)

This equation shows that, if the ambient concentration is null, the dilution factor defined in equation (1) and based on flux volumes also describes the dilution in terms of concentrations.

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The mass or volume dilution factor Sa is entirely defined by the density of the plume and, in principle, is independent on the contribution given by each individual pollutant. In recent years increasing emphasis has been devoted to the understanding of the concentration of pollutants in the ambient fluid at the end of the initial dilution phase. If both the effluent and the receiving fluids carry pollutant substances, the dilution reduces until in a limit it becomes zero when the concentration of pollutants in the environment equals or surpasses that of the effluent. On these grounds, for some purposes the ambient concentration cannot be taken as null, but the actual amount of substances in both fluids must be accounted for. In this case, the effective dilution factor (Saei) has to be used to measure the dilution of each pollutant in the plume (Baumgartner et al., 1994):

Saei 

cei  cai cpi  cai

(5)

where the suffix ―i‖ refers to the i-th substance and is used to underline that the effective dilution must be specified for each single substance. If cai = 0, equation (5) will be simplified to equation (4). Therefore, Saei measures the dilution of each individual substance rather than of the transport fluid. It differs from the average dilution (equation (4)) when the ambient fluid already contains a given amount of the pollutant (background pollution) (Frick et al., 2002). The background pollution acts as the minimum concentration that the plume can assume, and consequently sets an upper limit to the effective dilution.

Effluent dynamics The typical concentration profile of an effluent discharged in a fluid at rest is Gaussian (see, e.g., Csanady, 1980), i.e. highly diluted at the margins and concentrated in the centre (Figure 1). Such scheme is commonly used to represent the dynamics of a plume (Roberts, 1990; Skatun, 1996; MacIntyre et al., 1995; Glenn, 1997) and shows the formation of a minimum dilution zone in the central part where the highest concentrations are reported.

Plume concentration profile

Port diffuser

Figure 1. The dynamics of an effluent discharged in a fluid at rest from an horizontal diffuser. The figure shows the Gaussian profile of the concentration of the plume with high dilution at the margins and low dilution in the central part.

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Generally speaking, two kinds of plumes are considered (Frick et al., 2002): jet plumes and buoyant plumes. The jet plume (also referred to as pure jet) is a jet current with high momentum and no buoyancy, with a density similar to that of the ambient fluid at the discharge location. On the contrary, the buoyant plume (or pure plume) has null momentum and negative (i.e., the density of the plume is higher than that of the medium at the discharge location) or positive (i.e., the density of the plume is higher than that of the medium at the discharge location) buoyancy. The remaining types of plumes (defined as buoyant jets) have intermediate characteristics and may have at the same time momentum and buoyancy. Most of underwater discharges belong to this last category. Figure 2 shows a positively buoyant plume released by an horizontal diffuser located at a given depth of the water column. The effluent, lighter than the ambient fluid, is vertically curved and accelerated by its buoyancy. The region inside which the buoyancy and momentum of the discharge itself along with the effect of the local ambient current field results in rapid turbulent mixing of the plume, represents the near field.

depth

CURRENT

Diffuser

CENTERLINE

Plume diameter

Trap level

Plume maximum height

Horizontal distance Port elevation

Figure 2. The dynamics of a positively buoyant plume. The figure shows an effluent discharged from a diffuser under stratified water column conditions.

The entrainment process tends to decrease the density difference between the effluent and the medium. If the ambient fluid has constant density, the density of the diluted effluent will asymptotically reach the environmental density without equalling it, thus continuing its rise (or sink in the case of negative buoyancy), possibly reaching the surface (or the bottom). In stratified conditions, the effluent can reach a depth of the same density, called trap level or neutral buoyancy level. In the upper part of the trap level the ambient density is lower than that of the plume, whereas below it the density is higher; as a consequence, the effluent is trapped between these two layers. In general, when the effluent reaches the trap level it still has upward momentum bringing it in a region where the density is lower than that of the plume. For this reason, the buoyancy becomes negative, slowing down the plume until its vertical velocity is inverted and the plume becomes trapped. This second kind of trap level is different from that described above; during the rise and fall, the entrainment process still works determining changes in the

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average density of the plume. In an idealised case, the effluent oscillates around a variable trap level creating an oscillatory motion at the Brunt-Väisälä frequency (or buoyancy frequency). Once the initial kinetic energy is dissipated, the plume reaches a given depth and undergoes further transformations due to the ambient stratification and to the difference in density between the inner and the outer part of the plume. When neutral buoyancy is reached in the trap level, the density in the centre of the plume will be the same of the medium at the same depth. As mentioned above, the water column must be stably stratified (density increasing with the depth) to prevent any further vertical motion of the plume.

Densimetric froude number Most of the dilution occurs during the initial effluent discharge by entrainment. An important parameter to consider is the effluent densimetric Froude number (Fr), which evaluates the relationship between momentum and buoyancy of the plume as it is discharged through the port (Brandsma et al., 1992; Roberts and Tian, 2004):

Fr

U0 gD

(6)

a

e r

where: D = port diameter ρa = environmental density at discharge location ρe = effluent density ρr = reference density, usually equal to ρe g = gravity acceleration U0 is the velocity at which the plume is discharged through the port, and it is equal to:

U0

4q D2

(7)

where q is the flux volume at the port. If the diffuser is multi-port, then q=Q/n° ports, Q being the total flux volume. The quantity under the square root in eq. (6) must be positive. If ρa < ρe the absolute density difference is used for the square root, and only afterwards the minus sign is included in the equation. Fr is an adimensional number that allows the classification of the plume at the discharge location, differentiating between a jet plume and a buoyant plume; this parameter also measures the relative importance of buoyancy and momentum during the initial mixing. The numerator of eq. (6) is associated with the momentum of the plume, while the denominator represents the buoyancy. A low absolute Fr (around unity) indicates that the mixing process is dominated by buoyancy (buoyant plume), while higher values (usually greater than 10) are typical of

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momentum-driven mixing (jet plume). In addition, a positive or negative Froude number is respectively associated with a positively or negatively buoyant plume. A special case is verified when Fr is comprised between 0 and 1. In such a situation, the effluent is so buoyant that the environmental water might flow inside the port, thus causing the so-called ―saltwater intrusion‖. This effect might damage the port and alter its effectiveness. The saltwater intrusion mainly occurs in presence of horizontal ports, since the relevant positive buoyancy of the effluent determines an immediate rising of the plume, causing the inflow of seawater in the lower sector of the port. In order to avoid that this problem take place, a valve is usually installed to increase the Froude number. In offshore platforms PFW outfall pipes are generally installed in a vertical position. In empirical models a modified Fr, the Roberts Froude number (FRO), is commonly adopted, defined as (Roberts and Tian, 2004; MacIntyre et al., 1995):

Ua3

FRo g

a

e a

Q L

(8)

where Ua is the velocity of ambient current at port dept, while L is port length. The degree of initial dilution of an effluent is a function of environmental and technical factors. For example, a crucial parameter is the type of diffuser used to discharge the effluent (Roberts, 1990). The diffuser can be made of one or more ports, and even the diameter of each port can affect the velocity of the effluent and consequently its classification as jet or buoyant plume. The orientation of the diffuser can influence the plume trajectory, and in presence of a multi-port discharge also the distance between the ports can play a role. Another property affecting the entrainment and the dilution is the ambient velocity. Intense currents can increase the shear between the plume and the medium, thus enhancing turbulence, which in turn favours the entrainment of marine water within the effluent, and consequently its dilution. The stratification of the water column is another parameter influencing plume dynamics and dilution. In highly stable conditions, the plume can get trapped under the surface and the movements along the water column might be so reduced to prevent the dilution in the near field. The higher the rise (or sink) of the plume, the higher the chances for the effluent to mix with the medium. Under weak stratification (winter), an increase in the rise (or sink) depth promotes the dilution. In presence of strong stratification (summer) the situation is reversed (Petrenko et al., 1998).

CHEMICAL AND BIOLOGICAL OBSERVATIONS OF PFWS AT SEA In the past, the most relevant studies on PFWs have been carried out in areas characterized by the presence of offshore platforms discharging PFW into the sea (e.g. North Sea, Gulf of Mexico, Mediterranean Sea, coastal areas of Australia and of Southern California). These surveys generally have been designed to provide information on: a) chemical composition of PFW and fate of PFW contaminants in marine environment; b) toxicity of PFW and extent of toxic effects recorded in marine ecosystem; c) potential ecological impacts of PFW discharge on the marine environment.

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In the North Sea an important national monitoring programme was conducted by Norway to evaluate the fate of PFW in the marine environment and its effects on marine organisms. For these objectives, surveys were carried out from 1999 to 2003 to track PFW hydrocarbons in the environment and analyse the levels of these compounds in fishes living in Norwegian waters. Hydrocarbon levels in the marine environment were monitored using in situ natural and artificial sampling ―systems‖ such as mussels and semi permeable membrane devices (SPMD) located at different distances from the PFW discharge source. The results showed that even though hydrocarbons were recorded as far as 10 km from the discharge point, significant concentrations in terms of their possible biological effects were measured within a distance of 500 metres from the discharge. Some data on the potential impact of PFW originating from North Sea platforms are reported in the scientific literature (e.g. Somerville et al., 1987; Brendehaugh et al., 1992; Neff et al., 1992; Stagg et al., 1995; Stromgren et al., 1995; Frost et al., 1998; Henderson et al., 1999; Holdway, 2002). These studies show that the toxicity of PFW to marine organisms is low and would likely have acute effects only within the immediate mixing zone around production platforms. Effects of PFW may include altered benthic communities dominated by short-lived opportunistic polychaetes up to 100 m from offshore platforms. In the Netherlands the usefulness of the Chemical Hazard and Risk Management (CHARM) model to regulate use and discharge of chemicals employed in offshore exploration and production phases has been studied: trace metals and PAH concentrations have been analysed in mussels along a gradient from a PFW discharge. The results show that no increase was observed in mussel tissue in metal concentrations while a significant increase was measured for naphthalene (OGP, 2005). Bioaccumulation studies have been conducted in the Gulf of Mexico to evaluate PFW contaminant concentrations in edible tissue of fish and invertebrates collected near platforms (USEPA, 1993). The concentrations measured in this work resulted not harmful to fish, molluscs or to human health. However, Rabalais et al., (1992) observed bioaccumulation of hydrocarbons as far as 1000 metres from PFW diffusers and an ecological impact of PFW discharge in terms of the decrease of some benthic species near offshore platforms. Burns et al. (1999) used chemical tracers (benzene and toluene) to describe the distribution of PFW discharged from an offshore platform located in Northwest Shelf of Australia; they evaluated hydrocarbon bioaccumulation in bivalves in the vicinity of discharge point in connection with growth rates in natural marine bacterial and phytoplankton assemblages. The conclusion of chemical and biological analyses, associated to modelling studies, was that dissipation and degradation processes were fast and the area of potential biological impact extended as far as 900 m from the discharge. In a series of experiments described by Raimondi and Schmitt (1992), Osenberg et al. (1992), Krause (1993) and Krause et al. (1992) various invertebrates were exposed to PFW discharged from an outfall in the Santa Barbara Channel near Carpinteria (Southern California). Patterns of sub-lethal effects (in terms of reduced survivorship of larvae, settlement, ability to metamorphose, viability, ability of sperm to fertilize eggs and reproductive success) observed on different marine organisms (abalones, mussels and sea urchins) were inversely correlated with the distance from the diffuser. In the Adriatic Sea (Mediterranean basin) a monitoring programme is presently carried out in order to evaluate potential impact of PFW discharges. The monitoring plan includes chemical-physical analyses on seawater samples, chemical analyses on particulate matter

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within the water column, sediment and biota (mussels). Besides, biological investigations (ecotoxicological bioassays, biomarker analyses, study of benthic animal community and fish assemblage analysis) can fruitfully complement those analyses. Chemical composition and effects of PFW originated from Adriatic gas installations have been studied by several authors (Cicero et al., 2003; Mariani et al., 2004; Faraponova et al., 2007; Gorbi et al., 2007; Manfra et al. 2007; Fattorini et al., 2008; Gorbi et al., 2008). The chemical analyses show high concentrations of zinc and cadmium in sediment near the PFW discharge point. These contaminants are present in PFWs and they may be derived from corrosion or chipping of galvanized structures on the platform or in the oil/water separator system. Arsenic also showed higher quantities in biota near offshore platforms but chemical speciation allowed to exclude the anthropogenic impact connected with exploitation activities and revealed a natural regional gradient of arsenic levels (as arsenobetaine and arsenocholine) in mussel tissues. Different marine organisms (bacteria, algae, crustaceans, sea urchins and fish) exposed to PFW showed toxic responses but no significant toxic effect was observed when organisms were exposed to sea waters and to sediments collected along the PFW plume. The reason for this lies probably in the fact that these platforms drain only small volumes of PFW into the sea and the dilution process is rapid in the near field. Toxicity tests on PFW allowed to define a range of sensitivities for different test-organisms. Toxic effects of PFW on organisms may be due to absorption of water-soluble components through their surface epithelia and/or to oral ingestion and digestion of particulates. Besides, studies on benthic community show that densities of some benthic organisms (e.g. species ecologically linked to M. galloprovincialis) increase near offshore platforms (Trabucco et al., 2006).

NUMERICAL MODELLING OF PFW DISPERSION The complexity of the physical processes driving the dispersion of effluents discharged into the marine environment makes it difficult to assess the behaviour of the PFWs in the near- and far-field zones only by means of in situ observations. Relevant works (e.g., Petrenko et al.1998; Washburn et al., 1999; Nedwed et al., 2004; Cianelli et al., 2008) comparing model results with field measurements of effluent dilution showed that the numerical approach represents an efficient alternative to conducting difficult and expensive field observations in every discharge scenario. A numerical model may be defined as any tool describing a physical, chemical or biological problem by means of a set of equations solved by using different numerical methods. There is no ―best model‖, but rather a broad range of models that can be applied to simulate several different cases. The selection of a numerical model for a particular pollutant transport scenario requires considering various issues: Model aims: as a first natural step, the investigator has to define detailed model aims taking into account the physical processes involved in the system to be modelled. It is worth underlining that a model may target scientific as well as operational goals.  Model features: starting from the model goals, a list of the required model characteristics in terms of input-output flexibility has to be formulated. 

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Available models: the investigator may choose to apply an existing model or to develop a new one. Several models are available as public-domain, which allows to use widely tested and applied tools. If existing models are not suitable for a specific issue, a new model has to be devised and implemented.  Model choice: choosing a model requires an optimal compromise between the model aims, availability, cost/ease of use and accuracy. In the case of a new model, the cost and difficulty due to the implementation of the numerical tool depend on the available software and the complexity of the parameterized processes. The computational cost required to perform the calculation depends on the computer resources required (e.g., workstation, personal computer, etc.) and on the time and skills needed to run the simulation. The model accuracy may be determined comparing the model results with experimental data and is a fundamental step allowing to identify and estimate numerical errors. 

After selecting the model, a scale analysis is necessary to determine the relevant scales of the problem and the complexity of the model in order to adequately reproduce the studied system. A rigorous procedure to test a numerical model ensures that it is appropriate to simulate the functioning of our natural system. The following tests have to be carried out: verify the model consistency in terms of conserved quantities (e.g., mass);  test the model in idealized cases and compare the results with known analytical solutions;  calibrate and validate the model by comparing the numerical results with experimental data. 

A broad range of numerical models varying in complexity, accuracy and other features has been conceived for pollutant transport problems. It is worth underlining that accurate model implementation, calibration and testing are imperative to ensure the reliability of model results. The models may be classified into different groups according to the problem description adopted; transport processes can be described using the Eulerian or, equivalently, the Lagrangian approach. The difference lies in the expected output: the Eulerian approach will result in pollutant concentration maps, whereas the Lagrangian one will yield trajectories of pollutant particles. Following a general classification, the models simulating the dynamics of an effluent discharged into the fluid can also be divided into two main groups: empirical and theoretical models (MacIntyre et al., 1995; Glenn, 1997). The most frequently applied theoretical models are: the UM3 model (Three-dimensional Update Merge), the DKHW model (Davis, Kannberg, Hirst model for Windows) and the JETLAG (Lagrangian Jet) model (Baumgartner et al., 1994; Frick et al., 2002). Some of the most widespread empirical models are the CORMIX (Cornell Mixing Zone Expert System) and the RSB (Roberts-Snyder-Baumgartner) models (Baumgartner et al., 1994; Glenn, 1997; Frick et al., 2002). The better understanding of the physical processes involved in the pollutants transport problem has yielded a growing development of these advanced numerical models which accurately reproduce the dispersion processes and may provide an estimate of pollutants

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concentration generated by the effluent discharges. In particular, in the case of PFW releases, numerical modelling allows to simulate the dispersion process taking into account both the discharge and receiving environment conditions. For such reasons since the 1990s several transport models simulating the initial mixing process as well as the effect of the ambient currents and turbulence in the far- field zone on an effluent discharged into the sea have been successfully developed. Here we present a brief review of some of the most representative modelling studies on the dispersion of PFW into marine environment. Some of the previously cited numerical models have been successfully applied to evaluate the fate of PFWs discharged in coastal areas (e.g. Washburn et al, 1999; Berry, 2005; Cianelli et al, 2008). Washburn et al. (1999) used the RSB model to perform a field and modelling study around a diffuser located in California; they demonstrated that a crucial factor controlling the exposure of organisms to PFWs around the discharge point is the depth of the effluent in the water column. Berry (2005) developed an analysis of potential environmental effects associated with PFW discharge adopting an integrated modelling approach. In particular, the CORMIX model was applied to describe the dispersion of PFW released at Sable Island Bank (Canada); the results of this work suggested that the potential risks for the environment are low due to the rapid dilution of the wastewater plume. In the following section we will summarize the results of a case study (Cianelli et al., 2008) on the dispersion of PFW discharged in the Adriatic Sea (Mediterranean Sea), where the initial mixing has been simulated by means of the UM3 model. At present several modelling studies using various approaches have also been implemented and applied to PFWs discharged from platforms located in the main oil and gas extraction areas. In the North Sea, where the discharge of PFW from oil and gas production reached an annual volume of almost 400 million m3/year in 2003 (e.g. Durell et al., 2006), monitoring programmes have been conducted since the mid-1990s. These local and regional field studies have been used to optimize the monitoring plan as well as to implement and validate the numerical models simulating the dispersion and fate of the PFW chemical compounds released into the marine environment. The CHARM (Chemical Hazard Assessment and Risk Management) model (Stagg et al., 1996) was developed to predict the potential risks due to the chemicals released offshore and was validated with field measurements of concentration of selected PFW compounds. The DREAM (Dose related Risk and Effect Assessment Model) model was applied in the Norwegian sector of the North Sea to estimate the dispersion of PAH (polycyclic aromatic hydrocarbons) (Durell et al., 2006) and to predict the ecological risks associated with PFW discharges (Neff et al., 2006). A comparison with field measurements showed that the DREAM model results complement the in situ and laboratory data and that the numerical approach represents an useful tool for PFW discharges and impact assessment (Durell et al., 2006; Neff et al., 2006). The dispersion and dilution of PFW discharged from 95 oil platforms operating in the North Sea have also been simulated by means of the PROVANN (Produced Water in Norwegian) model (Rye et al., 1998). PROVANN is a three-dimensional model simulating the transport, dilution and degradation of chemical compounds released into the marine environment from one or more simultaneous discharge points (Reed et al., 1996). The

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numerical predicted concentrations were compared with the field data; the model provided useful results in terms of potential exposures to marine biota (Rye et al., 1998). Estimates of the PFW concentration in the Gulf of Mexico, the North Sea and the Bass Strait (Australia) were computed by means of the OOC (Offshore Operators Committee) Mud and Produced Water Discharge model (Brandsma and Smith, 1996). In a more recent work Smith et al. (2004) validated the OOC model using field data on mud and PFW discharges from platforms located respectively in California and in the Gulf of Mexico. In both studies, the model predicted plume depth and trajectory were in good agreement with field observations for a wide range of discharge and receiving environment conditions. In particular, in the near field zone the simulated PFW concentrations matched very accurately the measured data (Smith et al., 2004). Independently on the numerical approach followed, all the previously described works demonstrate that, at present, modelling PFW dispersion both in near- and far- field zones may play a crucial role in a ―prevention first― policy and represents an important first step in the design of a decision-making action.

MULTIDISCIPLINARY APPROACHES TO ASSESS THE FATE AND EFFECT OF PFW: THE CASE STUDY OF THE ADRIATIC SEA The impact of PFW discharges on the marine environment strictly depends on the characteristics of the effluent and the receiving ambient fluid. In order to assess the potential effects and risks for marine ecosystems it is of crucial importance to schedule monitoring plans. Monitoring programmes have been developed worldwide in all the areas (e.g. North Sea, Gulf of Mexico, Adriatic Sea) characterised by an intense extraction and production activity. OSPAR guidelines provide operational directions to plan the monitoring of the environmental impact of off-shore oil and gas activities. These monitoring programmes are mainly aimed at: a) establishing environmental conditions before the PFW discharge (background level); b) identifying spatial and temporal evolution of physical, chemical and biological parameters during the PFW discharge; c) assessing mitigation measures; d) defining guidelines supporting regulations and decisions. In order to study the effects of PFW discharge into the sea it is also very important to consider: a) which matrix has to be investigated (e.g. water column, sediment and biota); b) what sampling pattern and frequency have to be chosen; c) what parameters have to be measured. The most useful matrices in a PFW monitoring plan are: the water column since it receives the PFWs and because its physical-chemical characteristics influence the PFW

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dispersion; the sediment, which is a conservative matrix and a vehicle of transport for contaminants; finally, the biota providing information about ecological effects of discharges. Following the OSPAR guidelines, in a monitoring plan several sampling stations have to be scheduled. These stations need to be located taking into account the predicted extension of the area of influence as derived by discharge and local environmental conditions. During a monitoring plan several different parameters have to be measured (Maggi et al., 2006): -

physical (e.g. current speed and direction, temperature, salinity, density) chemical (e.g. oxygen and contaminant concentrations) geological (e.g. sediment grain size) geomorphological (e.g. bottom batimetry and morphology) ecological (e.g. benthic community) biological (e.g. bioavailability and toxicity of contaminants)

The comparison of these parameters with the background level can provide an indication on the potential levels of concern. For example, if the concentration of some substances in PFW is found to be significantly higher than the seawater background levels, then the bioavailability and ecotoxicity of these substances on selected species may have to be further considered. The toxicity studies are used as a complement to chemical measures for quantifying the potential toxic effects of PFW, together with the bioavailability of chemical substances and the possible synergies (see proceedings in Ray and Engelhardt, 1992 and in Reed and Johnsen, 1996; Manfra et al., 2007; Mariani et al., 2004). The in situ physical and chemical data may be also utilized in mathematical models to assess the time and spatial extent of PFW discharge effects at sea. For example, some chemical compounds of PFW (salts, nutrients, isotopes, ions) may be used as tracers. Their analysis in PFW and environmental samples (seawater and sediment) permits to track the PFW into the sea. A substance may be conveniently used as a tracer if it has some proprieties: conservative, representative of PFW, easy to analyse and to monitor. Burns et al. (1999) used benzene and toluene as tracers of PFW discharged into the North Sea. These compounds are volatile and so their concentrations in seawater often are very low. For this reason, they may be good tracers only when offshore platforms discharge high volumes of PFW, which permits to find detectable concentrations of BTEX into the sea. Cianelli et al. (2008) successfully used a chemical additive of PFW (DEG diethylene glycol) as tracer of PFW in the Adriatic Sea. Field measurements are often unpractical and expensive thus an exhaustive assessment of the effect of PFW discharge on marine environment needs to integrate in a monitoring plan field observations with results of numerical models of effluent dispersion. These models allow to perform realistic predictions of temporal and spatial extent of the PFW plume in selected regions as well as to easily evaluate several different discharge scenarios (dispersion under different environmental conditions or discharge characteristics). The application of multidisciplinary monitoring programmes allows a more complete assessment of the PFW fate and effects in the marine environment. Moreover, integrating field observations with a modelling approach may provide public administration and decision makers with useful directions to protect the marine ecosystem.

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A case study : the Adriatic sea (Eastern Mediterranean) In the last several years the offshore oil and gas extraction and production in the Mediterranean Sea has grown considerably. In particular, in the Adriatic Sea around 112 offshore platforms are currently active. The Adriatic is a shallow, landlocked semi-enclosed sea in the eastern Mediterranean, characterised by a wide shelf; with an average depth of few tens of meters in its northernmost section, and less than 200 in its middle portion, and a relevant seasonal variability in its circulation dynamics (e.g. Artegiani et al., 1997a-b). It is a basin characterized by strong gradients in water properties due to important surface (air–sea) and lateral (river runoff) fluxes; the dispersion of an effluent may vary on very short timescales due to current and water column stratification variability. In the Adriatic Sea thermohaline stratification and current field may also crucially affect the impact of PFWs discharges. Studies investigating the impact of offshore installations and PFW discharges in the Adriatic Sea have shown metal accumulation (Zn and As) in sediments and mussels (Mytilus galloprovincialis) very close to the platforms (Cicero et al., 2003; Manfra et al., 2007). As a case study of a multidisciplinary methodology applied to investigate the PFW dispersion, we here summarize the work by Cianelli et al. (2008). In particular in this work the authors integrated a numerical approach with a chemical one to investigate the near-field dispersion of the PFWs discharged from three offshore gas platforms located in the Northern Adriatic Sea. A numerical model was applied by Cianelli et al. (2008) in order to simulate the initial mixing of PFW during the summer and winter seasons and to analyse the dynamical factors influencing the dispersion of the plume in the marine environment. The numerical results were then compared with laboratory and field data of concentration of a chemical tracer observed in the receiving waters. In order to identify possible chemical tracers of the PFWs, the concentrations of the main groups of chemical compounds of PFW originating from the gas platforms were estimated. On the basis of preliminary analyses (Manfra, 2007), the chemical additive diethylene glycol (DEG) was chosen among the investigated chemical substances as the best affordable tracer of PFW. DEG is a chemical compound miscible in water and its presence in seawater may be exclusively associated with the PFW discharges, hence allowing to easily track the PFW plume (Weyerhaeuser, 2005). The authors used field data (PFWs and marine surface water samples) collected around the three gas platforms during summer 2005 and 2006 (Manfra, 2007) at increasing distance from the discharge source (0, 5, 10, 15, 20, 25 m from the diffuser). Temperature and conductivity measured during spring and summer seasons starting in April 2001 through August 2003 (ICRAM, 2002, 2003 and 2004) were also used. Hydrographic data (seasonal average stratification at the platform locations) for the winter season were derived from the Dartmouth Adriatic Data Base (http://thayer.dartmouth.edu/other/adriatic/databanks/hydrography/hydrography.html). Surface currents data (Table II) were drawn from Lagrangian drifter surface measurements conducted in the Adriatic Sea from 1994 to 1996 (Falco et al., 2000).

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D. Cianelli, L. Manfra, E. Zambianchi et al. Table II. Field data and input model values.

Platform

Average Port Water effluent diameter column flow depth (l d-1) (m) (m) 1 27400 0.45 116 2 11000 0.45 18 3 8200 0.45 23 * Outfall located above sea surface

Diffuser Depth (m) 3* 9 12

Surface current speed (cm s-1) 17 3 4

Effluent salinity

Effluent temperature

(g kg-1) 35 37.6 35.2

(°C) 23 20 20.5

Initial DEG conc. ( g l-1) 2040 9600 13000

The historical current data, the density profiles of the receiving water column, the measured DEG concentrations and the outfall pipe geometrical features represented the initial conditions of the numerical model. The near-field dispersion of the PFWs discharged from the three gas platforms was simulated by means of the UM3 (Three-dimensional Updated Merge) model (Baumgartner et al., 1994; Frick et al., 2002). UM3 is a three-dimensional Lagrangian theoretical model that quantifies the entrainment by applying both the Taylor entrainment (or shear) and the projected-area-entrainment (PAE) (or forced) hypotheses (e.g. Winiarski and Frick, 1976) (Figure 3). The Taylor entrainment is due to the shear between the effluent discharged and the receiving water, while the PAE entrainment is the rate at which mass is incorporated into the plume in presence of ambient current (Baumgartner et al., 1994). UM3 also computes the equations for conservation of mass, momentum and energy at each time step along the plume trajectory. The numerical output parameters (Table III) allowed to evaluate the space and time extent of the PFW plume. The results demonstrated a good agreement between the chemical analyses results and the numerical model outputs; both data sets highlighted that the PFW plume spread within a limited layer 2 to 4 meters thick, centered at the source depth (Figure 4). In low current conditions the initial dilution phase occurred within 15 m or less from the outfall. The results showed that also in the Adriatic Sea the PFW dispersion was mainly modulated by the seasonal stratification for the modelled current speeds. During the summer the trap level depth of the plume was deeper than in the winter season because of the high level of stability at the diffuser depth, while the low stratification occurring in winter conditions supported the vertical rise of the plume within the water column. v

v

a

b

Figure 3. Entrainment hypotheses applied by the UM3 model: a) shear entrainment and b) forced entrainment for a cylindrical fluid element.

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5

6

water column depth (m)

7

8

9

10 cln summer bnd summer cln winter bnd winter

11

12

13 0

2

4 6 8 plume horizontal distance from diffuser (m)

10

Figure 4. The vertical section of the plume vs the horizontal distance from the diffuser in summer (grey lines-closed circles) and winter (black lines-open circles) conditions. The solid lines represent the centerline of the plume while the dashed lines portray the plume boundaries (adapted from Cianelli et al., 2008).

Table III. Numerical model output parameters.

Platform

Summer DEG 1 2 3 Winter PFW 1 2 3

Length of initial mixing zone (m)

Trap level depth

Froude number

(m)

Effluent density

Ambient density

N Frequency

( t)

( t)

(rad s-1)

Dilution factor Sa

Initial mixing time (s)

15.0 2.3 14.8

Surface 9.8 11.5

0.16 0. 66 0.15

23.95 26.73 24.80

25.75 24.80 25.02

0.0758 0.3380 0.3326

175 55 74

90 70 400

10.0 6.8 6.7

Surface 7.8 10.8

0.09 0.06 0.03

23.95 26.73 24.80

29.00 27.78 26.50

0.0919 0.1539 0.2323

170 135 200

60 185 150

The modelled Froude number values (Table III) indicated that the water column stability at the platform locations influenced the PFWs dilution more than the local current field. In high stability conditions (summer) the simulations showed low values of the dilution factor Sa due to the limited extent of the initial mixing zone of the plume (Table III), whereas during

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winter the weak stratification sustained the dilution of the plume over a wide zone of the water column. In agreement with other works on PFW dispersion (e.g., Neff, 2002), the multidisciplinary approach applied in this case study in the Adriatic Sea also indicated rapid initial dispersion times which cause negligible or non toxic effects on marine organisms (Manfra et al., 2007). The integrated numerical-chemical approach applied by Cianelli et al. (2008) allowed to assess the dispersion processes influencing the potential effects induced on marine ecosystems by PFW discharges and provided suggestions to optimize the monitoring protocol presently adopted in the Adriatic Sea. In particular, on the basis of the results of this case study the monitoring plan in the Adriatic Sea was improved increasing the number and resolution of sampling stations in the horizontal and vertical around the discharge locations.

CONCLUSIONS The PFWs currently represent the largest waste stream derived from oil and gas offshore industry. The disposal of PFW in the marine environment presents the potential for a negative environmental impact. For this reason the regulatory environmental authorities of the countries involved in the extraction and production activities have established to continuously monitor the PFW discharges from offshore oil and gas platforms. In the last very few decades the comparison between laboratory and field observations and with numerical modelling results showed that acute effects may occur only within few hundred meters from the platform locations due to the rapid mixing and dilution of the PFW plumes. Consequently even the organisms entrained in the plume may be exposed in the worst cases to concentrations of pollutants decreasing on time scales of tens of minutes or few hours. In order to assess the fate and effect of the PFW in the marine environment the most promising monitoring plans have to integrate field observations and numerical modelling; such an approach may enhance our ability to understand and manage the potential effects of the extraction and production activities. The integrated monitoring approach is nowadays essential to support decisions in the assessment of the risk caused by the PFW discharges on marine ecosystems and should be available and commonly used by consultants, administrations and decision makers. An example is provided by the US EPA (Environmental Protection Agency) and by the Norwegian environmental authorities that usually include the dispersion modelling tools in the regulation of the offshore PFW discharges (e.g. Ray and Engelhardt, 1992). In the context of a cost-effective management of the marine environment, the disposal of PFWs is an important issue. The systematic application of integrated monitoring programmes may be very fruitful, as it can provide adequate indications to dispose the PFW properly, so as to help protect the marine environment while imposing at the same time the least economic burden to the oil and gas industry.

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Raimondi, P. T. & Schmitt, R. J. (1992). In Produced Water Technological/Environmental Issues and Solutions; J. P. Ray, & F. R. Engelhardt (Eds.), Plenum Press: New York, 415-430. Ray, J. P. & Engelhardt, R. (1992). Produced Water Technological/Environmental Issues and Solutions; Plenum Press: New York. Reed, M. & Johnsen, S. (1996). Produced Water 2: Environmental Issues and Mitigation Technologies (2nd ed.); Plenum Press: New York. Roberts, P. J. W. (1990). In The Sea: Ocean Engineering Science; B. Le Mehaute, & D. M. Hanes (Eds.), Wiley InterScience: New York, 661-689. Roberts, P. J. W. & Tian, X. (2004). Environ. Model. Soft., 19, 691-699. Rye, H., Reed, M., Melbye, A. & Johnsen, S. (1996). In Produced Water 2: Environmental Issues and Mitigation Technologies (2nd ed.); M. Reed, & S. Johnsen (Eds.), Plenum Press: New York, 258-269. Skatun, H. M. (1996). In Produced Water 2: Environmental Issues and Mitigation Technologies; M. Reed, & S. Johnsen (Eds.), Plenum Press: New York, 247-255. Smith, J. P., Brandsma, M. G. & Nedwed, T. J. (2004). Environ. Model. Soft., 19, 739-749. Somerville, H. J., Bennett, D., Davenport, J. N., Holt, M. S., Lynes, A., Mahieu, A., McCourt, B., Parker, J. G., Stephenson, R. R., Watkinson, R. J. & Wilkinson, T. G. (1987). Mar. Poll. Bull., 18, 549-558. Stagg, R., Gore, D. J., Whale, G. F., Kirby, M. F., Blackburn, M., Bifield, S., McIntosh, A. D., Vance, I., Flynn, S. A. & Foster, A. (1996). In Produced Water 2: Environmental Issues and Mitigation Technologies; M. Reed, & S. Johnsen (Eds.), Plenum Press: New York, 81-100. Strømgren, T., Sørstrøm, S. E., Schou, L., Kaarstad, I., Aunaas, T., Brakstad, O. G. & Johansen, Ø. (1995). Mar. Env. Res., 40(2), 147-169. Trabucco, B., Cicero, A. M., Gabellini, M., Virno-Lamberti, C., Di Mento, R., Bacci, T., Moltedo, G., Tommassetti, P., Panfili, M., Marusso, V. & Cornello, M. (2006). Biol. Mar. Medit., 13, 659-662. USEPA (1993), Federal Register Final MOD NPDES GOM Permit, Environmental Protection Agency, vol. 58, n. 231, 63964-63986. Utvik, T. I. R. (1999). Chemosphere, 39, 2593-2606. Washburn, L., Stone, S. & MacIntyre, S. (1999). Cont. Shelf Res., 19, 57-78. Weyerhaeuser. Water-Base Flexographic Ink, Material Safety Data Sheet MSDS., Weyerhaeuser Company, 2005. Number 135 - WC 379-01. Winiarski, L. D. & Frick, W. E. (1976). Cooling tower Plume model. USEPA Ecological Research Series, CERL, USEPA, Corvallis, EPA-600/3-76-100.

In: Fluid Waste Disposal Editor: Kay W. Canton, pp. 135-154

ISBN: 978-1-60741-915-0 © 2010 Nova Science Publishers, Inc.

Chapter 5

DISPOSAL OF SULFUR DIOXIDE GENERATED IN INDUSTRIES USING ECO-FRIENDLY BIOTECHNOLOGICAL PROCESS – A REVIEW A. Gangagni Rao* and P.N. Sarma Bioengineering and Environmental Centre, Indian Institute of Chemical Technology, Tarnaka, Hyderabad – 500807.

ABSTRACT Sulfur dioxide (SO2) is a known pollutant and responsible for various ill effects on living and non-living organisms. SO2 emissions can be reduced by using nonconventional energy sources or using conventional fuels containing less sulfur. However, under the present circumstances SO2 emissions cannot be completely avoided due to the reasons of rapid industrialization. Various technologies are available for the removal of SO2 from flue and waste gases. Most of these technologies fall under the category of physical, chemical or thermal. All these technologies generate secondary pollutants ending up in disposal problems and also cost prohibitive. Biotechnology offers relatively cheaper solutions for the conventional problems. Due to this reason, biotechnology is making in roads into the conventional treatment processes in all the fields. Over the last decade, efforts have been made to develop biotechnological alternatives to conventional physico- chemical processes for the removal of SO2 from flue gases. The only method available at present is Biological flue gas desulphurization (BIO-FGD).SO2 from flue gas can be absorbed in a suitable organic media. In the aqueous phase SO2 would be converted to sulfite and some part may again be converted to sulfate due to the presence of dissolved oxygen. Therefore, the aqueous phase will be having both sulfate and sulfite, which can be reduced to sulfide using Sulfate Reducing Bacteria (SRB) under anaerobic conditions. The sulfide formed in the anaerobic reactor could be converted to elemental sulfur using Sulfur Oxidizing Bacteria (SOB) under partial microbial aerobic conditions. The elemental sulfur can be used either as a soil conditioner or raw material for industrial applications. Therefore, BIO-FGD process could be an environmentally benign and economically viable alternative for the disposal of SO2 emitted from the industries *

Corresponding author: Telephone: +91-40-27191664, Fax: +91-40-27193159, E-mail: [email protected]

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A. Gangagni Rao and P.N. Sarma especially from power plants and refineries. The present article reviews the state of art of BIO-FGD process. .

Keywords: SO2, SRB, SOB, sulfate, sulfite, sulfide, sulfur, BIO-FGD.

INTRODUCTION Sulfur dioxide (SO2) represents the main fraction of anthropogenic sulfur emissions worldwide. Anthropogenic SO2 emission is mainly caused by combustion of sulfur containing fossil fuels such as coal and oil. Thermal based electrical power generating plants account for nearly 70% of all SO2 emissions. The flue gas contains SO2 in the range of 2000-4000 ppmv depending on the sulfur content of the coal or fuel that is being used. Another major source of SO2 is industrial combustion processes, metallurgical operations, roasting and sintering, coke oven plants, processing of titanium dioxide, pulp production and the thermal treatment of municipal and industrial wastes. Some non-combustion processes like production of sulfuric acid, treatment of metallic surfaces and oil refining processes (AIR Trends, 1995) also contribute to SO2 emissions. SO2 is a known pollutant and responsible for various ill effects for living and non-living organisms. (Koren,1991). It has a number of unwanted environmental effects like acid rain and formation of acidic aerosols. Therefore the need for SO2 removal from flue gases is evident and acknowledged by many countries (AIR Trends, 1995). Presently ambient air quality standards specify that SO2 concentration should not exceed 80, 60 and 15 μg/m3 in the industrial, residential and sensitive areas respectively. Many methods are available for controlling SO2 pollution in the industrial scenario. One of those methods is discharging into the atmosphere via stacks. Options for reduction of sulfur emissions at source include (Johns Eow, 2002; Rosenberg et al., 1975) measures of energy management, increasing the proportion of non-combustion renewable energy sources of the total energy supply, fuel switching (e.g. from high to low sulfur coals and or liquid fuels, or from coal to gas), fuel desulfurization and advanced combustion technologies (e.g. coal gasification combined with gas desulfurization). Another category of processes which aim at removing already formed SO2 is referred to as flue-gas desulfurization (FGD) (Pfeiffer, 1975). FGD is probably most widely used technique for control SO2 emissions from industries (DeNevers, 2000). State-ofart technologies for flue gas treatment processes are all based on the removal of SO2 by wet, dry or semi dry (also referred to as wet and dry absorption processes) and catalytic chemical processes. Biotechnology is being considered as an emerging technology in environmental protection, as it involves the use of microorganisms, which are more suitable for pollution control due to their versatility and minimize high chemical and catalyst costs to a certain extent. Apart from that microbiological processes operate around ambient temperature and atmospheric pressure. This eliminates the need of power utilization for heat and pressure and brings the energy costs down to the minimum. Disposal of by products of conventional physicochemical systems is another important drawback that may be overcome in a biological process. Over the last decade, efforts have been made to develop a biotechnological alternatives to conventional physico- chemical processes for the removal of SO2 from flue gas

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(Buisman and Prins, 1994). These biological methods for SO2 removal from flue gas are either direct or indirect. In direct biological methods, the feasibility of utilization of SO2 as an electron acceptor from flue gases by Desulfovibrio desulfuricans is studied (Lee and Sublette, 1991). Desulfotomaculum orientis grows on hydrogen (H2), carbon dioxide (CO2), and SO2 with production of hydrogen sulfide (H2S), while D. desulfuricans transforms SO2 to H2S using minerals and pretreated sludge as carbon source (Deshmane et al., 1993). Selvaraj et al. (1997a) carried out studies on various immobilized cell bioreactors to maximize the productivity of the bioreactor for SO2 reduction for effective flue gas desulfurization. The direct processes for SO2 removal through biological systems like biofilters suffer lot of limitations, such as the presence of high concentrations of CO2 (10–15%) in the flue gas, which can be inhibitory to the growth of the microorganisms in the biofilm. Further, direct biological processes are slow and it is hard to maintain a consistent supply of emissions at constant flow rate and at appropriate concentrations of the flue gas constituents in the bioreactor system. This problem can be solved by absorbing SOx selectively in a suitable solvent and the sulfate/sulfite-rich scrubbed solution can then be reduced to elemental sulfur using biotechnological approach. (Buisman et al., 1990a; Buisman and Prince, 1994; Cork et al., 1986; Janssen et al., 1995). In indirect biological processes, studies are carried out by using Thiobacillus ferrooxidans, for desulfurization of waste gas containing SO2 in a two step process. In this process, SO2 is scrubbed with ferric sulfate solution and the resultant ferrous sulfate solution is treated by Thiobacillus ferrooxidans bacteria and some are with sorbent regeneration process. Lee and Sublette (1991) proposed coupling the reduction of SO2 to H2S by mixed culture of sulfate- reducing bacteria containing D. desulfuricans to a Claus process as a means of by-product recovery from a dry generable scrubbing process of desulfurization of flue gas. Ligyphilip and Deshusses (2003) reported that complete treatment of SO2 from flue gases could be possible in a two-stage process consisting of a biotrickling filter followed by biological post-treatment unit. Buisman and Prins (1994) have proposed a new biotechnological process for flue gas desulfurization which consists of an alkaline SO2 absorption step followed by two biological steps. The first stage of the process is a chemical one in which flue gases containing SO2 are absorbed in a suitable solvent in the form of sulfate/sulfite/bisulfite and then biologically reduced to H2S. In the first biological step, sulfate is reduced to H2S, and in the second biological step H2S is converted into sulfur by colorless sulfur bacteria (Buisman et al., 1990a; 1991; Jassen et al., 1995). This process is cheaper than the presently used physicochemical processes and can remove up to 98% of SO2. Moreover, instead of gypsum or waste sorbent, this process produces sulfur that can be potentially be reused in the industry. It overcomes the drawbacks involved in direct biofiltration of SO2 and also includes the benefit of low-cost biological approach. In these biological processes the biological entities involved are sulfate reducing bacteria (Desulfovibrio, Desulfotomaculum, etc.), photosynthetic bacteria (genera of the families Chlorobiaceae and Chromatiaceae), and autotrophic Thiobacillus sp. A.G.Rao et al (2007) carried out work on microbial conversion of SO2 in flue gas to sulfide using bulk drug industry wastewater as an organic source by mixed cultures of sulfate reducing bacteria. In this study mixed cultures of sulfate reducing bacteria (SRB) are isolated from anaerobic cultures and are enriched with SRB media. Studies on batch and continuous reactors for the removal of SO2 with bulk drug industry wastewater as an organic source using isolated mixed cultures of SRB have revealed that isolation and enrichment methodology adopted in the present study are apt to suppress the undesirable growth of

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anaerobic bacteria other than SRB. Studies on anaerobic reactors showed that the process is sustainable at COD/S ratio of 2.2 and above with optimum sulfur loading rate (SLR) of 5.46 kg-S/m3/day, organic loading rate (OLR) of 12.63 kg COD/m3/day and at hydraulic residence time (HRT) of 8 hours. Free sulfide (FS) concentration in the range of 300 – 390 mg-FS/l is found to be inhibitory to mixed cultures of SRB used in the present studies. Bio -FGD occurs mainly in three steps (i) Conversion of SO2 to sulfite/ sulfate SO2 is absorbed either in water or in aqueous slurries of limestone and converted into sulfite and sulphate in a chemical scrubber (Janssen et al., 1995). SO2 + H2O Æ HSO3- + H+ HSO3-+1/2O2 Æ SO42 - + H+ The presence of oxygen in the flue gas results in the oxidation of part of the sulfite into sulfate. (ii) Biological conversion of sulfate/ sulfite into sulfide Sulfite and sulfate thus formed is reduced to sulfide (Widdel, 1988) by Sulfate Reducing Bacteria (SRB). HSO3- + 6 (H) + Æ HS-+3H2O SO42- + 8(H) + Æ HS-+3H2O+OH(iii) Biological conversion of sulfide into elemental sulfur Sulfide thus produced is partially oxidized to elemental sulfur by sulfide oxidizing bacteria (SOB) like Thiobacillus sp. under microaerophilic conditions (Buisman et al., 1990a; Buisman et al., 1991; Janssen et al., 1995; Anders B. Jensen and Colin Webb, 1995). HS-+1/2 O Æ Sº + OH-.

CONVERSION OF SO2 TO SULFITE/ SULFATE SO2 present in the flue gas reacts with water and forms the corresponding bisulfites. The presence of oxygen in the flue gas results in oxidizing part of the sulfite into sulfate. Main reactions which occur in the scrubber are as given below SO2 + H2O Æ HSO3- + H+ HSO3-+1/2O2 Æ SO42 - + H+ SO2 is absorbed either in water or in aqueous slurries of limestone and converted into sulfite and sulphate in a chemical scrubber (Janssen et al., 1995) which can be further converted into H2S and then into elemental sulfur by SRB and SOB respectively.

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Bandyopadhyay and Biswas (2006) studied the scrubbing of SO2 (initial concentration ranging between 400 and 1780 ppm) in a tapered bubble column scrubber using water and dilute sodium alkali. The results indicated that maximum water scrubbing efficiency of SO2 achieved in the tapered bubble column is 58.81 % / (W/m3) at a QL/QG ratio of 8.30 m3/1000 actual cubic meter (ACM). On the other hand, maximum water scrubbing efficiency of SO2 of 34.38 % /(W/m3) is reported at a QL/QG ratio of 10.17 m3/1000 ACM in a standard single stage bubble column where as almost 100% SO2 removal (i.e., zero penetration) is achieved in the scrubber developed in alkali scrubbing at an optimum QL/QG ratio of 3.0 m3/1000 ACM. Biological method involves passing of the flue gas through a biotrickling filter with Desulfovibrio desulfuricans. Under optimum conditions, this organism is shown to reduce SO2 to H2S within a contact time of 1-2 s. H2S thus produced is further converted to sulfate. A drawback in this approach is that D. desulfuricans is a strict anaerobe, and maintaining anaerobic conditions in biotrickling filters treating flue gases containing on an average 2-8% residual oxygen remains a challenge (Chou and Lin, 2000). Ligyphilip and Deshusses (2003) reported that complete treatment of SO2 from flue gases in a two-stage process consisting of a biotrickling filter followed by biological post-treatment unit is investigated. The biotrickling filter could remove 100% of influent SO2 from simulated flue gas at an empty bed residence time of 6 s for a concentration range of 300-1000 ppmv. All the absorbed SO2 is recovered in the biotrickling filter liquid effluent as sulfite (a product of chemical reaction of SO2) and sulfate (product of biological oxidation of sulfite). Biotrickling filter liquid effluent is further processed biologically in a single post-treatment unit consisting of a combined anaerobic and microaerophilic reactor for simultaneous reduction of sulfate and sulfite to sulfide and oxidation of sulfide to elemental sulfur. The post-treatment unit is used to treat effectively the biotrickling filter effluent and produce elemental sulfur. The sulfur production efficiency of the reactor is about 80% of the SO2 treated.

BIOLOGICAL CONVERSION OF SULFATE/ SULFITE INTO SULFIDE Biological sulfate reduction process is extensively studied and competes well with other sulfate removal technologies (Lens et al., 2000). Important aspects in the sulfate- reducing stage are the ability of the SRB to compete with other anaerobic bacteria for the available organic substrate and the sensitivity of the bacteria for sulfide. Addition of an appropriate electron donor is required for wastewaters that contain none or insufficient electron donor and carbon source for a complete sulfate reduction (van Houten et al., 1994; Warounsak Liamleam, Ajit P. Annachhatre, 2007). In BIO-FGD process, sulfite and sulfate, obtained by scrubbing SO2 from the flue gas with an alkaline solution, are anaerobically reduced to sulfide by SRB with an added electron H+ (Castro et al 2000, Hao et al 1994, Reis et al 1992, Widdle, 1988). A process is described whereby sulphate is reduced to sulphide under anaerobic conditions at a rate of 1.2 g SO4 1-1d-1 with producer gas (H2/CO) as substrate and 2.4 g SO4 1-1 d-1 with only CO as substrate. Stripped off sulfide is then converted to elemental sulphur by reacting it with biologically produced iron (III) (Dupreez and Maree, 1994). Sulfate reducing bioreactors for liquid waste treatment rely on biomass retention, as e.g. biofilters, UASB or gas lift reactors (Herrygers et al., 2000). An expanded granular sludge

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bed (EGSB) reactor, inoculated with acclimated sulfidogenic granular sludge, operated at 33 o C, fed with acetic acid as COD source and sulfate as electron acceptor, has a sulfate conversion efficiency of 80–90% at a high sulfate loading rate of 10.4 g SO42− S/l.d. The limitations of the EGSB technology with respect to the sulfate conversion rate appeared to be related to the biomass wash-out and deterioration of granules occurring at superficial upflow velocities above 10 m/h. Increasing the recirculation flow caused a drop in the sulfate reduction rate and efficiency (Dries et al., 1998). The gas lift reactor design using H2 as electron donor achieved so far the highest sulfate reduction rates, up to 30 g SO42− l−1 day−1 (van Houten et al., 1994, 1995b). Yet, this system is limited by mass transfer resistance of the gaseous substrate H2/CO2 to the biofilm and by the H2 lost when competition with methanogens and acetogens takes place. In a study carried out by Jan sipma et al (2007), it is observed that sulfate reduction rates are limited by the amount of CO supplied and its conversion efficiency (about 85%) at higher CO loads likely resulting from low biomass retention.

SULFATE/ SULFITE REDUCING BACTERIA (SRB) SRB can be defined as a mixed group of morphologically and nutritionally diverse anaerobic bacteria, which utilize sulfate as an acceptor for the dissimilation of organic compounds, and forms H2S as an end- product (Widdel, 1988). SRB are prokaryotic microorganisms, found in a variety of habitats (Odom, 1995). SRB may be spindle or beanshaped and the size is in the range of 1-10µm in length. Many of these bacteria do not show any distinctive colouration. They grow heterotrophically using small organic molecules, and autrophically using H2 as the electron donor and CO2 as the carbon source. SRB are capable of carrying out two types of sulfate reduction. Sulfate reduction to sulfide to meet only nutritional requirements of the bacteria is termed as assimilatory sulfate reduction. During this biosynthetic process, sulfide is not observed to accumulate in the environment. The other one called dissimilatory (or respiratory) sulfate reduction is characterized by the formation of large amounts of sulfide in the environment, which results in offensive odors, metal corrosion and failed digesters (Postgate, 1984; Lawrence et al., 1966; Anderson et al., 1982; Colleran et al., 1994). Of late, they have been used for the treatment of industrial wastewaters containing sulfate and heavy metals. Some typical examples are acid-mine drainage (Rowley et al., 1997), mineral processing- plant wastewaters (Barnes et al., 1991), tannery wastewaters (Shin et al., 1997), and gypsum wastes (Deswaef et al., 1996). In the recent past, various methodologies have been developed that allow rapid identification and quantification of bacteria in anaerobic bioreactors. Among those, the 16S rRNA hybridization technique is successfully used in many studies, and allows for detection and quantification of SRB in anaerobic biofilms and sludge (Amann, 1990, 1995; Bade et al., 2000; Deveruex et al., 1992; Loy et al., 2003; Purdy et al., 2001; Rabus et al., 1996; White and Gadd, 2000).

FACTORS INFLUENCING SULFATE/ SULFITE REDUCTION Compared with methanogens, sulphate-reducing microorganisms are very diverse in terms of their metabolic possibilities (Stams and Oude Elferink, 1997). The outcome of the competition of SRB with methanogenic bacteria (MB) for the substrates hydrogen and acetate

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and with acetogenic bacteria (AB) for intermediate substrates such as short-chain fatty acids (VFA) and alcohols is important. It will determine to what extent sulfide and methane, the end products of the anaerobic mineralization processes, will be produced. The following factors affect the biological reduction of sulfate and sulfite.

RATIO OF COD / SO4 2COD to sulphate ratio is a parameter widely used to control biological sulphate reduction when industrial effluent is used as sole organic carbon source. The COD/ SO4 2- ratio is the key factor in the partitioning of acetate utilization between MB and SRB. Importance of the competition between SRB and other bacteria (MB, AB) increases with decrease in the COD/sulfate ratio of the wastewater.

EFFECT OF PH & SULFIDE CONCENTRATION The most important conditions, which may play an important role in determining dominant SRB during digestion of sulfate-containing wastewaters, are the pH and the sulfide concentration of the bioreactor. SRB strains identified are sensitive to acidic waters (Hard et al., 1997). The acetate degrading SRB that are solar isolated, show an optimal pH in the range of 7.3 to 7.6. The minimum and maximum pH values for the growth are about 6.0 and 9.0, respectively (Sneath, 1984; Widdel and Pfennig, 1984; Widdel, 1988). In general, it is seen that the optimum pH values for the acetotrophic methanogenic bacteria (AMB) and acetotrophic sulfidogenic bacteria (ASRB) are in the same range.

EFFECT OF TEMPERATURE The sulphate reducing bacteria can be classified into mesophiles (growth temperature<40°C), moderate thermophiles (growth temperature: 40-60°C) and extreme thermophiles (growth temperature>60°C) based on their optimum growth temperature.

CONCENTRATION OF O2 SRB are strict anaerobic bacteria that are often in anoxic conditions in their natural biotopes. One of the most excessive biotopes where sulfate reduction coexists with anoxic conditions is provided in cyanobacterial mats. SRB are able to deal with temporary exposures to elevated oxygen concentrations up to 1.5mM (Sigalevich et al., 2000).

EFFECT OF CONCENTRATION OF METAL IONS Many heavy metals are toxic to microorganisms such as SRB and influence the sulfate reducing capacity of SRB due to their capacity to deactivate enzymes by reacting with their

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functional groups, to denature proteins, and to compete with essential cations. Utgikar et al. (2001) reported that the insoluble metal sulphide formed is not toxic to SRB by itself but it blocks the access to substrate and the nutrients that are essential for bacteria by forming a precipitate coating the SRB.

ELECTRON DONORS Waste gas scrubbing waters originating from flue gas desulphurization units present a special problem, since they do not contain organic compounds to support the SRB. In order to biologically treat these waste streams, an external carbon and energy source has to be supplied. The choice for the appropriate electron donor is based on the suitability for the sulfate reduction process, and the availability in large enough quantities at a cheaper or no cost. In any potential application of SRB bioreactors, the selection of electron donor source is expected to have an impact on the economics of the process and will be a site-specific design criterion.

BIOLOGICAL CONVERSION OF SULFIDE INTO ELEMENTAL SULFUR SOB produce elemental sulfur as an intermediate in the oxidation of H2S to sulfate. Sulfur produced by these microorganisms can be stored in the sulfur globules, located either inside or outside of the cell. The excreted sulfur globules are colloidal particles. The nature of sulfur and surface properties of the globules are different for sulfur produced by different bacteria. The sulfur produced by phototrophic bacteria appears to consist of long sulfur chains terminated with organic groups where as chemotrophic bacteria produce globules consisting of sulfur rings. The partial oxidation of H2S to elemental sulfur (S0) instead of sulfate (SO42-) has several advantages. When compared to H2S, elemental sulfur is non-toxic, non-corrosive solid containing more sulfur per unit mass. The main use of elemental sulfur is as a feedstock for the chemical, fertilizer and material manufacturing industries. Elemental sulfur is a desired end product of sulfide oxidation because of the nature of settleability, less oxygen requirement and possibility of its reclamation and reuse as a valuable byproduct in industrial applications and metal bioleaching processes even though oxidation of sulfide to sulfate yield more energy for the bacteria than in the formation of sulfur. In the biotechnological process, dissolved sulfide (HS-) is converted to elemental sulfur. The insoluble sulfur can easily be removed from the water stream and can be reused as a soil fertilizer or fungicide. Biological conversion of sulfide into elemental sulfur by using SOB is shown schematically in Fig.1 (University of Copenhagen, 2006).

Fig. 1: Biological conversion of Sulfide into Elemental sulfur by SOB

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Cork et al; (Cork et al., 1983; Cork, 1982a; Cork and Garunas, 1982b; Cork and Ma, 1982c) proposed a microbial process for the removal of H2S from a gas stream based on the photosynthetic bacterium Chlorobium thiosulfatophilum as an alternative to the Claus or Stetford process. In nature sulfide can be oxidized biologically in three different ways (Kuenen, 1975): 1. Anaerobic oxidation by photosynthetic bacteria. 2. Oxidation by denitrifying organisms. 3. Oxidation with oxygen by the colorless sulfur bacteria. Removal of sulphide is brought about by photoautotrophic bacteria in suspended reactors at sulphide loading rates up to 100 mg/ h.l. Chemoautotrophs also successfully removed sulphide at higher loading rates. Gangagni Rao et al., (2003) developed a stripper, which is capable of removing sulfide (60 to 70%) from anaerobically treated wastewater, before aerobic treatment, without altering the chemical characteristics of the wastewater. It is a physical system in which air and wastewater are passed as counter currents. H2S may be efficiently removed by coupling this type of stripper to the existing anaerobic systems. The treated wastewater from the stripper, which contains fewer sulfides, may be post-treated in the aerobic system before final discharge. Ravichandra et al., (2006) studied fluidized bed bioreactor to test the biological conversion of sulfides using immobilized cells of Thiobacillus sp., that are isolated from aerobic sludge of distillery and dairy effluent treatment plant using standard methods. Henshaw et al. (1998) reported that complete removal of S2- from the influent in a continuous stirred-tank photosynthetic bioreactor could be achieved with more than 90% conversion of the removed sulfide to elemental sulfur. In the case of chemotropic reactors (Buisman et al., 1990a), cent per cent of the sulfide removed from the effluent is converted to elemental sulfur. Kobayashi et al., (1983) proposed the use of photosynthetic bacteria i.e. chlorobium thiosulfatophilum for sulfide removal where sulfur is the end product. However the requirement of radiant energy for photosynthentic bacteria is an economic disadvantage. Gommers et al., (1987) and Sublette and Sylvester (1987a, 1987b) investigated the use of denitrifying bacteria for sulfide oxidation and sulfur as the end product. This system is not widely applicable because nitrate is additionally needed as oxidizing agent. The most important bioconversion in an aerobic sulfide-oxidizing bioreactor is (Kuenen, 1975): 2 HS- + O2 Æ 2 S0 + 2 OHStudies conducted on biological sulfide oxidation using suspended and submerged attached growth systems revealed that biological sulfide oxidation is somewhat limited due to difficulty of maintaining adequate biomass level in the former case whereas uneven oxygen profile in the bioreactor led to the formation of more sulfate and thiosulfate in the latter case. Moreover, in the submerged attached growth system, anaerobic condition often prevailed within the inner part of the biofilm that led to bioconversion of sulfur to sulfide again by heterotrophic sulfur reducing bacteria. Both biomass limiting conditions and uneven oxygen distribution could be eliminated by employing biotrickling filter. Sulfide oxidation strongly depends upon oxygen concentration, sulfide loading rate and upflow velocity (Annachhatre and Suktrakoolvait, 2001). Anaerobic conversion of the

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sulfur particles is obviously detrimental to the efficiency of the system because it leads to an increased consumption of the required electron donor and increased sulfide levels in the anaerobic reactor, which may cause inhibition of the biomass (Koster et al., 1986).

SULFIDE OXIDIZING BACTERIA The desirable bacterium for the bioprocess is that which will be able to convert H2S to elemental sulfur. It should require a minimum of nutrient input and produce elemental sulfur that is easily separable from the biomass. Most of the SOB is spherical or rod-shaped organisms. SOB is often embedded in a gelatinous mass and range from less than 1µm to several tens or hundreds of µm in length. All contain a type of bacteriochlorophyll, a pigment related to the chlorophyll found in plants, algae and cyanobacteria. Cell masses may be yellow-green-brown, purple-violet-red, or green in colour. SOB live in special tissues in invertebrates such as Riftia pachyptila (vestimentiferan tube worms) and Calyptogena magnifica ('giant' white clams) that live around deep sea hydrothermal vents. Here they provide energy, by oxidizing reduced sulphur compounds, and organic matter, by converting carbon dioxide to organic compounds, which the invertebrates use. Metabolism for sulfur bacteria is shown in Fig.2 (University of Copenhagen, 2006)

Fig.2: Metabolism for sulfur bacteria (University of Copenhagen, 2006)

Recognized groups of SOB are mainly phototrophic bacteria and chemotrophic bacteria. Phototrophic bacteria (green or purple in colour) use light as energy source to reduce CO2 to carbohydrates. Reduced sulfur compounds are used as electron donors for this reduction. The reduction takes place under anaerobic conditions. Phototrophic C. limicola is an ideal bacterium in these biological removal processes due to its ability to grow under anaerobic conditions using only inorganic substrates and a light source and its efficient extracellular production of elemental sulfur from H2S. The chemotrophic sulfur bacteria (colourless bacteria) obtain energy from the chemical aerobic oxidation of reduced sulfur compounds i.e. spontaneous reaction of H2S or elemental sulfur with dissolved oxygen at the water surface. SOB can also be classified into lithotrophic bacteria, which use inorganic substances as source for hydrogen and organotrophic bacteria, which use organic substances as source for hydrogen. The process for sulfide removal is based on aerobic oxidation by the group of colorless sulfur bacteria (Kuenen and Beudeker, 1982). To this group belong the organisms

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with widely different types of physiology and morphology. Genera belonging to the group of colorless sulfur bacteria are: Thiobacillus, Thiomicrospira, Thermothrix, Thiothrix, Thiospira Pseudomonas, Thioovulum, Sulfolobus, Beggiatoa, and Thioploca. In Thiobacillus sp., Thiobacillus ferrooxidans, Thiobacillus thiooxidans, Thiobacillus novellas, Thiobacillus thioparus, Thiobacillus denitrificans etc. are major sulfide reducing bacteria. Visser et al. (1997a) isolated the dominant autotrophic sulfide oxidising strain present in elemental sulfur producing bioreactors, which is found to be a new Thiobacillus species, designated as Thiobacillus sp. W5. This organism is then used as a model organism for studying sulfur production by thiobacilli in wastewater treatment reactors (Visser et al., 1997b). Interestingly, the end product of sulfide oxidation at a given sulfide loading rate by Thiobacillus sp. W5 is almost exclusively elemental sulfur. A very closely related bacterium, Thiobacillus neapolitanus, converted only 50% of the sulfide to elemental sulfur while the other 50% is completely oxidised to sulfate. Comparison of the metabolic properties of Thiobacillus sp. W5 with those of Thiobacillus neapolitanus revealed that Thiobacillus sp. W5 has a competitive advantage over Thiobacillus neapolitanus in bioreactor environments because its sulfide oxidising capacity is double that of Thiobacillus neapolitanus. Interestingly, the maximum specific oxygen uptake rates of the two organisms are very similar. No other significant biochemical differences are observed between the two organisms. This means that the limited sulfide oxidation rate of Thiobacillus neapolitanus gives this species a competitive disadvantage, as it can oxidise only 50% of the incoming sulfide to elemental sulfur. Thus, these bioreactor environments select species such as Thiobacillus sp. W5.

FACTORS EFFECTING SULFIDE OXIDATION Effect of Oxygen Rate The sulfide is biologically oxidized to elemental sulfur under oxygen-limiting conditions (Buisman and Prins, 1994). Kuenen (1975) has postulated that the existence of sulphur produced enhances its chance of further oxidation to sulphate or formation of linear polymer, polysulphide respectively. Therefore the sulphur produced should be simultaneously removed from the reactor. 2 HS- + 4 O2 Æ 2 SO42- + 2 H+ n/8 S8 + HS- Æ

S2-n+1 + H+

Previous studies by Janssen et al. (1995) have demonstrated that the molar oxygen to sulphide ratio for sulphur production would be around 0.6 –1.0. However, in practical situation such as a real wastewater treatment plant, it is difficult to maintain a narrow sulphide/oxygen ratio. On the other hand maintaining optimum redox potential could control sulphide oxidation more precisely (Janssen et al., 1998; Khanal and Huang, 2003). At reactor dissolved oxygen (DO) concentrations higher than 0.1 mg l-1, sulfate is the main product of

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sulfide oxidation. At DO less than 0.1 mg l-1, sulfur is the major end product of the sulfide oxidation.

Effect of pH Sulfur compounds, when biologically degraded in gasphase bioreactors, generate sulfate leading to a substantial drop in pH with concomitant reduction in biological activity and thus in H2S removal. Different species are active in the pH range 0.5 to 10 (Bergey, 1974). Most of the previously reported biological sulphide oxidation (BSO) systems are operated around neutral pH in the range of 7–8 (Janssen et al. 1995, 1997; Hasan et al., 1994; Lee and Sublette, 1993; Buisman et al., 1990b, 1991; 1989; Sublette, 1987). Janssen et al. (1998) have achieved maximum sulphur recovery in suspended growth system maintained at pH 8. Yang and Allen (1994a, b) observed removal efficiencies exceeding 99% under a wide range of heterotrophic conditions in a compost biofilter. However, a rapid decrease of pH led to reduced removal efficiency. Cox and Deshusses (2002) observed that operating a bioreactor at either pH 4.5 or 7.0 has not affected the performance of the biotrickling filter significantly, when feeding H2S and toluene are fed simultaneously. However, they have found that at pH 4.5, the start-up phase for toluene degradation is relatively long and that sudden pH drop may cause temporary poor removal of H2S and toluene. The H2S removal efficiency of biotic increased as the pH of the nutrient feed increased in the pH range of 2.0–6.0. However, an opposite trend is observed for the pH between 6.0 and 7.0. The optimum pH for the autotrophic population is near 6.0. Krishna Kumar et al., (2005) reported that up to 19 kg sulphide/m3d loading, the sulphide removal is nearly 100%, in which sulphur recovery is around 80% and sulphate is around 2– 3% of total sulphide sulfur when its pH is controlled at 8 in Reverse Fluidized Loop Reactor (RFLR). In the case of chemotrophic SOB such as the Thiobacillus denitrificans in the RFLR, the optimum pH for the growth and sulphide oxidation is around 7–8 (Janssen et al. 1995; Buisman et al., 1989; Sublette, 1987; Krishnakumar et al., 2005; Sublette and Sylvester, 1987a). Under the same sulphide loading, the biomass experienced more toxicity at pH 9 and 9.5 compared to the pH at 8. The presence of higher residual sulphide and transitional products such as S2O32- and polysulphide are a clear indication of decline in SOB activity in RFLR (Krishnakumar et al., 2005) under alkaline pH. However, employing alkaliphilic SOB, as reported recently (Banciu et al., 2004; Sorokin et al., 2004) in future bioreactor studies may help to remove sulphide under more extreme conditions.

Effect of Temperature SOB exists in environments with temperatures up to 100 oC (Tributsch, 2003). There is little difference in the amount of sulfide oxidation at either ambient room temperature or 35 o C. Increasing the biooxidation temperature from 35 to 50 oC increased apparent sulfide oxidation from about 38% to 48%. At 60 oC, sulfide oxidation is highest at about 51% (James Brierley, 2003).

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SUMMARY Biotechnological methods for the removal of SO2 gas from flue gas streams is a promising alternative when compared to the other available physical and chemical methods especially flue gas streams having low SO2 concentrations. This process is also applicable for the removal of SO2 in tail gases emitted from processes treating high concentrations of SO2. The biotechnological methods involve exploitation of microbial metabolism to convert the toxic SO2 compound into inoccus elemental sulfur. The BIO-FGD method is a three step process; 1. Absorption of SO2 gas and conversion of the gas into sulfites and sulfates. 2. Anaerobic treatment of sulfite and sulfate containing solution by suitable microbial consortia for conversion into sulfides 3. Aerobic oxidation of sulfides to convert into elemental sulfur by microorganisms. Conversion of SO2 into sulfites and sulfates is a physical and chemical process which is not elaborated in the present review. Extensive discussions are done for anaerobic treatment of sulfites and sulfates by sulfate reducing bacteria. Metabolism of SRB requires organic carbon as carbon source and sulfates and sulfites as energy sources. Proper selection of suitable organic sources and the ratio of COD/SO4 2- play a major role in efficient removal of sulfates and inhibition of methanogenesis. The anaerobic sulfate reducing bacteria are sensitive to variation in environmental conditions such as pH, temperature, presence of O2 and presence of metal ions. Controlling of all the parameters and optimization is crucial for the efficient operation of the sulfate reducing reactor. In the biotechnological process, dissolved sulfide (HS-) is converted to elemental sulphur by the aerobic metabolism of SOB. The insoluble sulfur can easily be removed from the water stream. Conversion of sulphide to elemental sulphur is a sensitive step and is dependent on dissolved oxygen concentration. Temperature, sulphide loading rate and pH are the other critical parameters in aerobic sulphide oxidation process. Optimization of parameters in all stages of BIO-FGD process is crucial for efficient removal of SO2 gas from flue gas streams.

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In: Fluid Waste Disposal Editor: Kay W. Canton, pp. 153-181

ISBN: 978-1-60741-915-0 © 2010 Nova Science Publishers, Inc.

Chapter 6

NOVEL BIOLOGICAL NITROGEN-REMOVAL PROCESSES: APPLICATIONS AND PERSPECTIVES J.L. Campos, J.R. Vázquez-Padín, M. Figueroa, C. Fajardo, A. Mosquera-Corral and R. Méndez Department of Chemical Engineering. School of Engineering. Rua Lope Gómez de Marzoa s/n. University of Santiago de Compostela, E-15782, Santiago de Compostela, Spain.

ABSTRACT Since the requirement for nutrient removal is becoming increasingly stringent, a high efficiency of nitrogen removal is necessary to achieve a low total nitrogen concentration in the effluent. Biological nitrification and denitrification processes are generally employed to remove nitrogen from wastewater. Unfortunately, these processes are not suitable to treat wastewater with a low COD/N ratio because it involves the addition of an external organic carbon source and, therefore, an increase of the operational costs. Several alternative processes for nitrogen removal can be applied in order to reduce partially (―nitrite route‖) or totally (anammox, autotrophic denitrification) the organic matter required. Such processes suppose not only an economical way to treat these wastewaters but they are also more environmentally friendly technologies (lower production of CO2, N2O and sludge; lower energy consumption). Up to now, they were basically applied to the return sludge line of municipal wastewater treatment plants (WWTPs). However, these processes could even be implemented in the actual WWTPs in order to achieve more compact and energy efficient systems. Their potential advantages can make them also feasible technologies to treat polluted ground water or to remove nitrogen compounds from recirculating aquaculture systems.

Keywords: Anammox, aquaculture systems, autotrophic denitrification, BABE, CANON, COD/N ratio, ground water, nitrite, SHARON, WWTPs.

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1. INTRODUCTION The main reserve of nitrogen is located in the atmosphere, where it is found as N2, but this molecule can not be directly used by most living beings. Both bacteria and blue-green algae can carry out nitrogen fixation converting N2 into ammonia and nitrate. Once nitrogen has entered the biosphere via biological fixation, it is subject to a series of conversion steps, from plant protein to animal protein, and finally ends up in dead organic matter. When this organic matter is mineralised in the soil, most of the inorganic nitrogen compounds produced (ammonium, nitrite, nitrate) are taken up again by plant or microbial biomass for the production of protein. The nitrogen cycle also includes electrochemical reactions (electrical storms) although their importance is small when compared to biological processes (Gijzen, 2001). Human activity is changing the rate of some processes involved in the natural nitrogen cycle which causes accumulation of nitrogenous compounds in water and, therefore, its associated pollution. The most harmful anthropogenic processes are: a) Mining and the use of fossil fuels since inactive nitrogenous compounds are introduced again into the active cycle and b) fixation of nitrogen from the atmosphere by chemical processes (fertilizer industries) and by intensive cultivation of nitrogen fixation plants. The problems generated by the nitrogenous compounds in wastewaters will depend on their oxidation state: Ammonia can have toxic effects on aquatic life; these effects may be either acute (i.e., fish mortality) or chronic (impacts on reproduction, tumours, etc.). The presence of ammonia also causes algal blooms (eutrophization) which influence the water system in two ways. First, they hamper the penetration of sunlight, causing death of underwater grasses. Secondly, the decomposition of dead algae causes depletion of oxygen, which is normally essential to most organisms living in water. Nitrite can bind to iron on hemoglobin reducing transfer of oxygen to cell tissues; the result is suffocation accompanied by a bluish tinge of the skin. Nitrite and nitrate in drinking water have been medically linked to methemoglobinemia, a sometimes fatal blood disorder affecting infants (―blue baby syndrome‖). On the other hand, the disposal of wastewater containing nitrite and/or nitrate can generate both NO and N2O by an incomplete denitrification. These compounds contribute to the destruction of the ozone layer. Nitrate pollution impedes the production of drinking water. During chlorination of drinking water, carcinogenic nitrosamines may be formed by the interaction of nitrite with compounds containing organic nitrogen. To avoid these problems, legislation has imposed maximum disposal limits for nitrogenous compounds. These limits can be achieved by physico-chemical or biological treatments, the last one being the most used for economical reasons.

2. BIOLOGICAL NITROGEN REMOVAL PROCESSES Nitrogen appears in nature under different oxidation states. The main biological transformations between the existing nitrogen compounds are fixation, ammonification,

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assimilation, denitrification and anammox, which are carried out by different microorganisms (Figure 1). The COD/N ratio of wastewater will determine which of these biological processes is the most suitable to remove nitrogen: -

COD/N >20: Removal by nitrogen assimilation by heterotrophic bacteria. 20>COD/N>5: Removal by assimilation, nitrification and denitrification. COD/N 5: Removal by partial nitrification-denitrification or partial nitrificationanammox.

2.1. Nitrification-Denitrification The combination of nitrification-denitrification processes is generally applied to remove nitrogen from municipal wastewater and most industrial wastewaters. Nitrification is a twostep process: ammonia is firstly oxidized into nitrite and then nitrite into nitrate (Equations 1 and 2). These steps are carried out by autotrophic ammonia- and nitrite-oxidizing bacteria, respectively (Khin and Annachhatre, 2004). NO2- + H2O + 2 H+

NH4+ + 1.5 O2

NO2- + 0.5 O2

[1]

NO3-

[2]

During denitrification both nitrate and nitrite are reduced to nitrogen gas under anoxic conditions, organic matter being used as electron donor (Equation 3). This process is carried out by denitrifying bacteria.

m Deco

n io Ni

De

ni

Vegetal protein

at

N org

tio n

Fix

NO2-

ica

ing feed

o ec

on ati Chemical industry

m

D

tio

N2

Atmospheric N2 Nitrogen

fic

n

si po

im

Assim il

ation

Figure 1. Nitrogen cycle.

on

al Anim

s As

tio

ati

n

ila

ox

Animal protein

mm

Urine Urea

N org

a An

s

i lis

o dr Hy

fic

Faecal matter

Fixation

o

NH+4+

NH4 Ammonium

ion posit

ri Nit

m Am

n

ic nif

tri f

o ati

8HCO3- + 6H2O + 2CO2 + 4N2

tr i

8NO3- + 5CH3COOH

NO3-

[3]

158

J.L. Campos, J.R. Vázquez-Padín, M. Figueroa et al. Predenitrification Wastewater COD NH4+

N2

O2

NH4+ Denitrification

Nitrification

Effluent

COD

NO3-

Postdenitrification Wastewater COD NH4+

N2

O2 NO3Nitrification

Denitrification

Effluent

COD

Figure 2. Schematic representation of predenitrifying and postdenitrifying configurations.

Integration of both nitrification and denitrification also allows reducing the amount of chemicals needed to control pH during the treatment since alkalinity generated during denitrification compensates for the pH decrease due to nitrification. Nitrification occurs under aerobic conditions while anoxic conditions are necessary for denitrification. This implies the use of two different tanks which can be provided in two possible configurations (Figure 2): a) Predenitrifying configuration—wastewater is fed into the denitrifying reactor and later nitrification is carried out. A stream from the aerobic tank containing nitrate and/or nitrite is recirculated to the first unit to carry out denitrification. Therefore, nitrogen removal efficiency depends on the recycling ratio; b) Postdenitrification configuration—wastewater is fed into the nitrifying unit and its effluent enters into the denitrifying reactor. This configuration is very simple, easy to control and no recycling is needed. Nevertheless, organic matter is oxidized in the aerobic unit and an external carbon source must be added for denitrification which increases the operational costs. This configuration is only used when the COD/N ratio of wastewater is low. In the predenitrification configuration, the organic matter coming from the denitrification unit causes the proliferation of heterotrophic bacteria in the aerobic unit. These microorganisms compete with nitrifying bacteria for oxygen and its concentration must be maintained at levels around 1–2 mg O2/L to avoid the failure of the nitrification process. On the other hand, the concentration of nitrifiers in the aerobic tank is low due to their slow growth rate and, therefore, the required volume of the aerobic unit is high. The use of carrier material in this unit would increase the concentration of nitrifiers and, then, decrease the required volume [Pegasus system (Tanaka et al., 1996)].

2.2. Partial Nitrification Nitrification and denitrification processes are suitable to remove ammonia from wastewater when its COD/N ratio is high but operational costs increase when no organic matter is available (for example effluents from sludge digesters, landfill leachates, effluents from anaerobic digesters of fish canneries…).

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To treat such effluents a postdenitrification configuration should be used, being necessary the addition of an external carbon source (methanol, acetic acid…) to complete denitrification. In those cases, partial oxidation of ammonia into nitrite would suppose a decrease of both oxygen and organic matter requirements of 25% and 40%, respectively, and only 60% of sludge is generated compared to the full oxidation into nitrate (Van Kempen et al., 2001) (Figure 3). Under normal operational conditions of municipal wastewater treatment plants (pH= 6.58.5; T= 10-20 ºC; dissolved oxygen= 1-2 mg O2/L; low ammonia concentrations), no nitrite production is observed. Since the nitrification process consists on two serial reactions, nitrite accumulation will occur when the rate of ammonia oxidation is higher than that of nitrite oxidation. To achieve this aim, factors affecting to both reactions must be changed in such a way that the growth rate of ammonia-oxidizers is more enhanced than that of nitrite-oxidizers. Both rates depend on temperature, dissolved oxygen and free ammonia concentrations (Park and Bae, 2009) (Equation 4): max

(T)

O2

[4]

S

KO2 O2 K S (1 NH3 ) S KINH3

where μmax is the maximum growth rate (d-1), O2 the dissolved oxygen concentration (mg O2/L), KO2 the oxygen affinity constant (mg O2/L), S the substrate concentration (NH4+ for ammonia-oxidizers and NO2- for nitrite-oxidizers) (mg N/L), KS the substrate affinity constant (mg N/L), NH3 the free ammonia concentration which inhibits both ammonia- and nitrite oxidizers (mg N/L) and KINH3 the free ammonia inhibition constant (mg N/L). 1 mol Nitrate (NO3-)

40%Organic matter

Nitrifiers

Heterotrophs

Aerobic-Nitrification

Anoxic-denitrification

25% O2 1 mol Nitrite (NO2-)

1 mol Nitrite (NO2-)

60% Organic matter

75% O2 1 mol Ammonia (NH4+)

Nitrification-denitrification 4.6 g O2/g NH4+-N oxidized 7.5 g DQO/g NO3--N reduced

½ mol Nitrogen Gas (N2)

Partial nitrification-denitrification Advantages; 25% Reduction of O2 demand 40% Reduction of required organic matter 40% Reduction of biomass generated

Figure 3. Conventional nitrification-denitrification process compared to partial nitrificationdenitrification process.

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J.L. Campos, J.R. Vázquez-Padín, M. Figueroa et al. Table 1. Kinetic parameters of ammonia- and nitrite oxidizers (Wiesmann, 1994). Parameter KO2 (mg O2/L) KINH3 (mg N/L) Ea (kJ/mol)

Ammoniaoxidizers 0.3 116

Nitrite-oxidizers

68

44

1.1 0.52

Table 1 shows the values of the different kinetic parameters. Since the oxygen affinity constant of nitrite-oxidizers is higher than that of ammonia-oxidizers, a decrease of the dissolved oxygen in the reactor would exert a higher effect on the former one. Therefore, nitrite generation would be favoured at low dissolved oxygen concentrations (Figure 4a). However, this operational strategy implies also the decrease of the ammonia-oxidizers activity and a control system for oxygen is required when the influent characteristics are not constant. It is also known that ammonia oxidation is inhibited by higher free ammonia concentrations than nitrite oxidation. Therefore, partial nitrification could be achieved by maintaining free ammonia levels in the reactor which only causes the inhibition of nitriteoxidizers (Figure 4b). Albeit, the maintaining of a certain concentration of free ammonia in the system means that the effluent does not fulfil the disposal requirements. Another disadvantage of this strategy is the possible adaptation of nitrite-oxidizers to free ammonia and the restoration of their activity. Since the activation energy of ammonia oxidation is higher than that of nitrite oxidation, an increase of temperature will cause a higher effect on the first step. In practise, the ammonia oxidation rate is higher than the nitrite oxidation rate at temperatures higher than 30 ºC (Figure 4c). From an economic point of view, this strategy would be only feasible when effluents have already such temperature, for example effluents of anaerobic digesters.

2.3. Anammox Process Anammox process (Anaerobic Ammonium Oxidation) was discovered around 15 years ago in the Technical University of Delft (The Netherlands) during the operation of a denitrifying pilot plant treating wastewater of a yeast company. This process is carried out by a group of autotrophic bacteria capable of oxidizing ammonia into nitrogen gas using nitrite as electron donor. Neither oxygen nor organic matter are needed in this process (Equation 5): NH4+ + 1.32 NO2- + 0.066 HCO3- + 0.13 H+ 1.02 N2 + 0.26 NO3- + 0.066 CH2O0.5N0.15 + 2 H2O

[5]

These bacteria belong to the phylum Planctomycetes and their yield coefficient is low (Y= 0.038 g VSS/g NH4+-N). This low sludge production reduces the management costs but makes the start-up of anammox reactors quite long. For this reason, systems with high biomass retention capacity must be used. Another characteristic of these microorganisms is

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the decrease of their activity in the presence of oxygen, nitrite or organic matter (DapenaMora et al., 2004; 2007). Growth rate (d -1)

1

A

ammonia oxidizers

0.8 0.6 0.4

nitrite oxidizers

0.2 0 0

1

2

3

4

O2 [mg/L]

Growth rate (d -1)

1.2

B

ammonia oxidizers

1 0.8 0.6 0.4

nitrite oxidizers

0.2 0 0

0.2

0.4

0.6

0.8

1

NH3 [mg N/L]

Growth rate (d -1)

3

C

2.5

ammonia oxidizers

2 1.5 1

nitrite oxidizers

0.5 0 10

15

20

25

30

35

40

T [ºC]

Figure 4. Possible strategies to carry out partial nitrification by changing dissolved oxygen concentration (A), free ammonia concentration (B) and temperature (C).

To apply the anammox process, the effluent should contain suitable concentrations of both ammonia and nitrite. Ammonia is generally present in wastewater but nitrite is not and has therefore to be generated by the oxidation of 50% of the ammonia. During partial nitrification, organic matter is also oxidized which prevents its possible negative effects on the anammox reactor. The combination of anammox and partial nitrification to treat wastewaters with high nitrogen content and without organic matter gives some advantages compared to the conventional nitrification-denitrification process (Figure 5): 1) The oxygen

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requirements are 60% lower; 2) No organic matter must be added; and 3) Sludge production is 85% lower (Fux and Siegrist, 2004). Nitrification/denitrification

Partial nitrification/anammox

1 mol NH4+ CO2

1 mol NH4+ CO2

1.9 mol O2 (100%) 3.4 g CODbiomass 1 mol NO357 g COD

17 g CODbiomass 0.5 mol N2 + 20 g CODbiomass

0.4 mol O2 (40%) 1.6 g CODbiomass 0.55 mol NO2-

0.45 mol NH4+

CO2

1.5 g CODbiomass 0.45 mol N2 + 0.1 mol NO3-+ 3 g CODbiomass

Figure 5. Comparison between nitrification-denitrification and partial nitrification-anammox processes to treat wastewaters with low COD/N ratios (Adapted from Fux and Siegrist, 2004).

2.4. CANON Process Under limiting oxygen conditions (lower than 0.5% of air saturation) a mixed culture of both ammonia-oxidizers and anammox bacteria can be obtained. This culture converts ammonia directly into nitrogen gas with nitrite as intermediate product. Nitrifiers consume oxygen and generate both nitrite and an anoxic environment for anammox microorganisms. Then, ammonia can be removed in a single unit under autotrophic conditions. Different acronyms were used to define this process: OLAND (Oxygen-Limited Aerobic Nitrification and Denitrification) (Windey et al., 2005), aerobic deammonification (Wett, 2006) and CANON (Completely Autotrophic Nitrogen removal Over Nitrite) (Sliekers et al., 2002; 2003). The two former names are based on the idea that the own ammonia-oxidizers carried out the denitrification process. However, nowadays, it was demonstrated that anammox bacteria are responsible of the denitrification process, the last acronym being the most suitable to define the process. Two possible strategies to start-up a CANON system are possible: 1) to inoculate an anammox reactor with nitrifying biomass and to supply air to maintain microaerobic conditions or 2) to operate a nitrifying reactor under oxygen limited conditions to obtain the desired ammonia to nitrite molar ratio inside the system and then to inoculate anammox biomass (Pynaert et al., 2004; Gong et al., 2007). The second strategy seems to be more suitable because an important decrease of the anammox activity is observed when the first strategy is applied (Sliekers et al., 2002; 2003; Liu et al., 2008). Moreover, only a few amount of anammox biomass is necessary to start-up the CANON process with this second strategy (Vázquez-Padín et al., 2009a; 2009b).

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2.5. Autotrophic Denitrification Sulfur compounds can be present in wastewaters together with carbon and nitrogen compounds and the interactions between the biological cycles of the three elements can be used to remove them (Figure 6). The biological interaction between sulfur and nitrogen cycles is given by autotrophic denitrification which consists on the reduction of nitrogen oxides (NO3- and/or NO2-) into nitrogen gas by using reduced sulfur compounds as electron donors (S2O3-2, Sº and/or H2S) (Equations 6, 7 and 8). The end product is sulfate which is less harmful than nitrate, especially when the effluent is disposed in a marine environment. 5 S2O3-2 + 8 NO3- + H2O 5 S-2 + 8 NO3- + 8 H+ 5 S° + 6 NO3- + 2 H2O

10 SO4-2 + 4 N2 + 2 H+

[6]

5 SO4-2 + 4 N2 + 4 H2O

[7]

5 SO4-2 + 3 N2 + 4 H+

[8]

Autotrophic denitrification presents the following advantages compared to heterotrophic denitrification: a) No carbon source is required; b) Elemental sulfur can be used as an economical electron source; c) The sludge production is lower (Campos et al., 2008). The main autotrophic denitrifying bacteria belong to the Thiobacillus denitrificans and Thiomicrospira denitrificans genus. These microorganisms are mainly mesophilic with an optimum temperature of 25-35 °C while their optimum pH is 7-8. Contrary to heterotrophic denitrification, this process requires a source of alkalinity (HCO3-, CaCO3) to neutralize the protons produced (Fajardo et al., 2008).

Figure 6. Biological interactions between carbon, nitrogen and sulfur cycles.

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Figure 7. Configuration of a WWTP with sludge treatment.

These bacteria can also use oxygen as electron acceptor to oxidize sulfur compounds. The end product is sulfate under high dissolved oxygen levels while, at low oxygen concentrations ( 0.1 mg O2/L), only a partial oxidation into S° occurs (Equations 9 and 10). If the aim of the treatment is to remove nitrate, the presence of oxygen must be avoided since the microorganims will preferentially use it as electron acceptor. 2HS- + 0.5O2 2HS- + 4O2

S°

+ H2O

[9]

2SO4-2 + 2H+

[10]

3. NITROGEN REMOVAL TECHNOLOGIES BABE, SHARON and partial nitrification-anammox technologies, which will be described in the following sections, have already been applied at industrial scale to treat sludge digesters effluents of municipal WWTPs (Figure 7). Nitrogen concentration of these effluents ranges between 300 and 1,700 mg N/L and contributes to 15-20% of the total inlet nitrogen load of the WWTP although its flow rate is only about 1% (van Loosdrecht and Salem, 2006). As this stream is previously treated by the anaerobic digester, it contains a low amount of organic matter and its temperature is around 30 °C. On the other hand, its HCO3/NH4+ ratio is 1, that is, the available alkalinity only allows nitrifying 50% of ammonia.

3.1. BABE (Bio Augmentation Batch Enhance) Bioaugmentation of nitrifying bacteria in the flocculent sludge could reduce the required volume of the aerobic tank. This may be achieved by promoting nitrification of the sludge digester effluent with biomass coming from the return sludge stream. Only part of this stream

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should be applied to the effluent in order to maintain its temperature around 20-30 °C. This process has been successfully applied at industrial scale and was named BABE (Salem et al., 2002a; Berends et al., 2005). Another advantage of its application is the increase of the denitrification capacity of the WWTP since part of the volume of the aerobic tank is not required and can be operated under anoxic conditions. The BABE process is operated with denitrification in order to control pH. In this aspect, the sludge supplied to the process allows minimizing the amount of external organic matter added. Under anoxic conditions, the electrons needed to reduce nitrate are given by endogenous respiration of sludge. The BABE process can be carried out in a system with one or two units. The system with one only unit is operated in cycles. Firstly, the reject water and sludge are fed to the system to nitrify under aerobic conditions. During a second stage, denitrification occurs and sludge partially settles. At the end of this stage, the liquid fraction and the non settled sludge are fed to the main stream of the WWTP. The two reactors configuration consists of an anoxic reactor following by an aerobic tank (Figure 8). In the anoxic tank, the supernatant of the sludge digester is mixed with the sludge which also acts as carbon source and even external organic matter could be added if necessary. A recirculation between both units is maintained in order to supply nitrate to the anoxic tank. The BABE process with a single unit has been tested at industrial scale in the WWTP of Garmerwolde (The Netherlands) with a capacity of 300,000 inhabitants-equivalent (Salem et al., 2004). The implementation of this process allowed improving ammonia concentration in the effluent from 13.3 to 5.2 mg NH4+-N/L. In order to upgrade the WWTP of Walcheren to fulfil disposal requirements ( 10 mg N/L) (140,000 inhabitants-equivalent, The Netherlands) two alternatives were evaluated: An increase of both anoxic and aerobic tanks and the implementation of the BABE technology. The second option allowed reducing 50% of the required area and supposed saving costs of 115,000 Euros/year (Salem et al., 2002b) (Table 2). This technology has also been implemented in the WWTP of Hertogenbosch (350,000 inhabitants-equivalent, The Netherlands).

Figure 8. Configuration of the two units BABE process.

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Table 2. Comparison of the upgrading of the WWTP of Walcheren (The Netherlands) by the conventional procedure and by bioaugmentation in order to fulfil disposal requirements ( 10 mg N/L) (Salem et al., 2002b). Item

Upgrading by the conventional method

Aerobic tank Anoxic tank BABE system Total increase

Upgrading by the implementation of the BABE process Increase (%) (Additional volume required/initial volume) 88 22 1,300 370 0 14 225

75

3.2. SHARON (Single-Reactor System for High Activity Ammonia Removal Over Nitrite) The SHARON process takes advantage of the higher growth rate of ammonia-oxidizers compared to nitrite-oxidizers at high temperatures (Hellinga et al., 1998; van Dongen et al., 2001; Mosquera-Corral et al., 2005). This process is carried out in a conventional CSTR with suspended biomass, without sludge retention which operates at elevated temperatures (30–40 °C) and at a HRT of 1 day to promote the wash-out of nitrite-oxidizers while ammoniaoxidizers are retained (Figure 9). This process is combined with denitrification, by adding methanol, to control the pH and to remove the nitrite generated. Both nitrification and denitrification can take place in the same unit with an intermittent aeration or in two different units. Table 3. SHARON reactors in operation (van Loosdrecht and Salem, 2006). Plant

Capacity (inhab-eq)

Utrecht Rotterdam Zwolle Beverwijk Garmerwolde Den Haag New York

400,000 470,000 150,000 320,000 300,000 1.100,000 3.000,000

Nitrogen load (kg N/d) 900 830 540 1.200 700 1.200 5.500

Year of start-up 1997 1999 2000 2004 2004 2005 2007

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Figure 9. One single unit SHARON process.

The SHARON process was developed in 1997 and the first plant at full scale was built the same year. Nowadays, full-scale sludge liquor treatment with partial nitrification/denitrification in SHARON reactors has already been introduced in 7 WWTPs (van Loosdrecht and Salem, 2006) (Table 3). Economic balances demonstrated the cost savings using the combined SHARON-denitrification processes to treat reject water in comparison to physicochemical processes or conventional nitrification-denitrification processes (Table 4). The costs distribution of this treatment are: 47% installation costs, 15% energy costs, 4% maintenance, 8% working costs, 18% methanol and 7% management of sludge produced. Table 4. Estimation of costs for nitrogen removal in the sludge line for a WWTP of 500,000 inhabitants-equivalent (van Kempen et al., 2001).

Physico-chemical process Stripping with air Stripping with steam Precipitation Nitrification/denitrificatio n Membrane reactor Airlift reactor SHARON/denitrification

Chemical sludge

Biologica l sludge

Energy requiremen t

Cost (Euro/kg N)

Yes Yes Yes

No No No

Normal High Low

6.0 8.0 6.0

No No No

Yes Low Low

High Normal Normal

2.8 5.7 1.5

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3.3. SHARON-Anammox The combination of partial nitrification and anammox processes allow minimizing oxygen and organic matter requirements during the sludge liquor treatment. This technology can be applied in systems of one and two units. In the two-units configuration, a SHARON reactor is used to nitrify only 50% of ammonia by controlling both dissolved oxygen concentration and pH. Later, anammox process is carried out in UASB (Upflow Anaerobic Sludge Blanket) o IC (Internal Circulation) reactor similar to those used during anaerobic treatment at high loading rates (Figure 10). Activated sludge reactor Anoxic

Aerobic

Effluent

Influent

Sludge return SHARON reactor

Secondary settler Thickening tank

Anammox reactor

Water line Sludge line

Sludge digester Dehyrated sludge

Dehydration system

Figure 10. SHARON-Anammox process.

In the one-unit system, partial nitrification and anammox processes are simultaneously carried out under microaerobic conditions. For this propose systems with flocculent ammonia-oxidizing and anammox granular biomasses can be used. Another option is the utilization of granular biomass where ammonia-oxidizing bacteria grow in the outer layers consuming oxygen and generating nitrite and, therefore, suitable conditions are promoted in the inner layers to develop the anammox process (Figure 11). Under a practical point of view, systems of one unit are preferred because higher removal rates can be achieved (smaller reactors) and their N2O emissions are low although systems with two units are more flexible and stable against influent fluctuations (Kampschreur et al., 2008).

Novel Biological Nitrogen-Removal Processes: Applications and Perspectives

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PARTIAL + NITRIFICATION NH4 O2

ANÓXICA AEROBIC NO2-

ANOXIC N2 + NO3-

NO2NH4+

ANAMMOX

Figure 11. Simultaneous partial nitrification and anammox processes in granular systems.

Up to now, there are 4 anammox plants operating at full scale (Abma et al., 2007), three of them in The Netherlands and one in Japan (Table 5). All of them have reached their design capacity treating wastewater from different origins which indicates the wide applicability of the process. It is important to point out that the length of the first reactor (72 m3) start-up was 3 years while the fourth plant was started up in 2 months (Van der Star et al., 2007). This fact was mainly due to a higher knowledge about anammox process and a greater availability of inoculum. Table 5. Full-scale anammox plants around the world (Abma et al., 2007).

Project

Application

Design capacity

Load achieved

(kg N/d)

(kg N/d)

Start up time

Waterboard Hollandse Delta, The Netherlands

Municipal

(2 units)

(reject water)

490

750

3.5 years

Tannery

325

1501

1 year

Potato processing

1200

7001

6 months

Semiconductor

220

220

2 months

IWL, The Netherlands (2 units) Waterstromen, The Netherlands (1 unit) Mie prefecture, Japan (2 units) 1

No more nitrogen available.

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Under the denomination of deammonification (DEMON), there are other two full scale plants in Austria (500 m3, 300 kg N/d) and Switzerland (400 m3, 250 kg N/d) treating the supernatant of sludge digesters (Wett, 2006; 2007). The start up length of the first plant was 2.5 years while the plant located in Switzerland, inoculated with sludge from the first one, got its design load in only 50 days. Economic estimations were calculated with data obtained from these full scale plants and compared to those obtained when the nitrification-denitrification process is used. The results showed important economic and environmental benefits (Table 6). Table 6. Comparison between traditional treatment and partial nitrification-anammox process. Nitrificationdenitrification

1

Energy (kWh/kg N) Methanol (kg/kg N) Sludge production (kg VSS/kg N) CO2 emissions (kg/kg N) Total costs1 (Euros/kg N)

Capital and operational costs included.

2.8 3 0.5-1.0 4.7 3-5

Partial nitrificationAnammox 1 0 0.1 0.7 1-2

The application of this process to reject water would suppose important saving costs (Siegrist et al., 2008). Most of the municipal WWTPs designed only for organic matter removal were equipped with primary settlers with a hydraulic retention time (HRT) of 2–3 h to reduce the organic matter applied to the biological reactor. Albeit the requirement of nitrogen removal caused that the HRT of primary settlers was reduced to less than 1 h in order to ensure the availability of organic matter during denitrification. This fact implied the decrease of biogas generated during sludge anaerobic digestion. If the partial nitrificationanammox process is applied to treat reject water, the denitrification capacity of the WWTP could be decreased without negative effects on the overall nitrogen removal efficiency. This would allow increasing again the HRT of the primary settler to enhance production of biogas and significantly decrease the requirement of oxygen in the aerobic tank to remove organic matter. On basis of a nitrogen removal efficiency of 75%, the treatment of reject water would allow decreasing a 25% the denitrifying capacity. Then, this process would need 25% less organic matter which can be separated in the primary settler by adding flocculant increasing the biogas production in 25% (Figure 12). Taking into account that oxygen requirements for organic matter removal and nitrification suppose 70-80% of the total energetic costs of the plant, this new configuration could achieve a reduction up to 50% of the energy consumed (Table 7).

Novel Biological Nitrogen-Removal Processes: Applications and Perspectives Nitrogen balance

171

Denitrification Primary effluent 6,0 g N/p·d 4,5 g N/p·d

10,5 g N/p·d 8,5 g N/p·d

Inlet

Biological treatment

Primary settler

Outlet

a 10 g N/p·d b 10 g N/p·d

2,5 g N/p·d 2,5 g N/p·d Primary sludge 1,0 g N/p·d 1,5 g N/p·d

1,5 g N/p·d 0 g N/p·d

Anammox

Supernatant Digester supernatant

2,0 g N/p·d 1,5 g N/p·d Digested sludge Sludge digester

1,5 g N/p·d 1,5 g N/p·d

1,5 g N/p·d 1,5 g N/p·d

1,5 g N/p·d

Degraded

COD balance

Primary effluent 85 g DQO/p·d 65 g DQO/p·d

Inlet

Secondary sludge

Primary settler

40 g DQO/p·d 30 g DQO/p·d Biological treatment

Outlet

a 110 g DQO/p·d b 110 g DQO/p·d

5 g DQO/p·d 5 g DQO/p·d Secondary sludge

Primary sludge 25 g DQO/p·d 45 g DQO/p·d

40 g DQO/p·d 30 g DQO/p·d

Digested sludge

Biogas 30 g DQO/p·d 40 g DQO/p·d

Sludge digester

35 g DQO/p·d 35 g DQO/p·d

Figure 12. N and COD balances of a municipal WWTP with two possible configurations of the primary settler: a) HRT: 0.5-1 h; b) HRT: 2 h, addition of flocculant and treatment of reject water with the partial nitrification-anammox process (Adapted from Siegrist et al., 2008).

Table 7. Estimation of the energy net consumption based on mass flows presented in Figure 12 (Siegrist et al., 2008). Mass flow (g/p·d) Case a Case b Electric energy consumption -COD removal -Nitrogen removal Electrical power for pumping and mixing Electrical power generated from biogas Net energy consumption

40 22

30 22

30

40

Energy (kW/p·d) Case a Case b 0.040 0.022 0.020 0.038 0.044

0.030 0.022 0.020 0.051 0.021

Comparison of the available technologies to treat rejected water The bioaugmentation, partial-denitrification and partial nitrification-anammox technologies are applied to optimize the nitrogen removal in WWTP (van Loosdrecht and Salem, 2006). The selection of one of the strategies will depend on the specific limitations in the nitrogen removal of each plant according to the decision tree represented in Figure 13. The first decision concerns whether the limiting step in the nitrogen removal in the main plant

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is the nitrification or the denitrification process. In case nitrification or denitrification is limited by the volume of the aerobic or the anoxic tank, respectively, the bioaugmentation is the recommended technology. However, if the aeration or the organic matter is limiting the nitrogen removal, the ―nitrite route‖ processes are recommended. WWTP must reach full nitrification in order to apply these processes since the selective increase of ammonia-oxidizers could cause the presence of nitrite in the final effluent. The ammonium counter-ion plays an important role and will determine the most suitable technology to be applied. In most WWTP this counter ion will be the bicarbonate ion which works as pH buffer in the nitrification process and the recommended technology would be the partial nitrification-Anammox. Finally, if the sludge is dried the counter-ion will be the acetate ion; in this case, the partial nitrification-denitrification will be the most suitable technology. Besides these aspects, the start-up time, the risk of failure, the flexibility of the process, etc. will also determine the decision of what technology to implement.

3.4. ANANOX (Anaerobic-Anoxic-Oxic) The anaerobic treatment of effluents containing sulfate (canneries, petrochemical industries, tanneries, etc.) implies generation of sulfide. This compound presents some problems such as: a) odour problems and toxicity; b) decrease of organic matter removal efficiency and, therefore, less methane generated; c) corrosion problems; d) the need for biogas conditioning and postreatment of effluents. A simple method to remove sulfide is autotrophic denitrification. Nowadays, only one plant is applying this process at full scale which is called ANANOX (Figure 14). This plant is located in Italy and treats municipal wastewater (Garuti et al., 2001). Supernatant

Nitrification space

Nitrification Limiting factor

Limiting process

Denitrification

Limiting factor

BABE

Aeration capacity

Denitrification space

Organic matter

Counter-ion NH4+

Acetate

HCO3-

SHARON

SHARON ANAMMOX

Figure 13. Diagram of selection of the best technology to treat reject water.

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NO3-

Influent

Activated sludge reactor

Sludge tramp

Anaerobic

Settler Effluent

Anoxic

Thickening tank

Water line Sludge line

Sludge purge

Figure 14. ANANOX process.

This technology is based on a two units configuration. The first unit is an anaerobic reactor with three compartments (2 anaerobic + 1 anoxic) containing flocculant sludge. The second unit is a conventional activated sludge system with a settler. In the first unit, anaerobic digestion of organic matter and sulfate reduction into sulfide are carried out. During autotrophic denitrification, sulfide is again oxidized into sulfate with the nitrate coming from the effluent recirculated. Activated sludge reactor Influent

Primary settler

Anaerobic

Anoxic

Aerobic

Effluent

Secondary settler

Thickening tank Secondary sludge

Thickening tank Primary sludge

Water line

Thermal hydrolysis reactor

Dehydration system

Sludge line

Biogas Dehydrated sludge Sludge digester Crystallization unit

CANON system

P recovery

Figure 15. Application of sludge reduction techniques in WWTPs.

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4. APPLICATION PERSPECTIVES 4.1. Municipal WWTPs Improvement Excess sludge treatment and disposal of conventional WWTPs supposes between 50 and 60% of operational costs. For this reason, recently, a great effort in the development of new technologies to reduce sludge production was done (Kroiss, 2004). When a sludge digester is already present in the WWTP, the implementation of a sludge disintegration unit (for example, thermal hydrolysis) previous to the anaerobic digester is the best option to maximize the recovery of energy from the sludge (Figure 15). This treatment would allow an increase in methane production and would decrease the HRT of the sludge digester. Ammonia concentration in the digester supernatant would increase, the application of a CANON system would be even more profitable from an economical point of view. This system could be operated to obtain an effluent with a stoichiometric NH4+/PO4-3 ratio to obtain struvite (Equation 11). The phosphorus recovered can be used as fertilizer. Mg+2 + NH4+ + HPO4-2 + OH- + 5H2O

MgNH4PO4· 6H2O

[11]

4.2. New Concept of WWTPs Anaerobic digestion could be an interesting alternative to the aerobic process to treat domestic wastewater since it is a net energy producing process with a lower sludge production (1/10 ratio compared to aerobic systems). Since the rate of the anaerobic processes strongly decreases at temperatures lower than 20 C, systems with a good biomass retention capacity, such as the upflow anaerobic sludge blanket (UASB) reactor or expanded granular sludge bed (EGSB) reactor are needed to make anaerobic treatment of municipal wastewater a feasible option. Nevertheless, the low hydraulic retention time applied to these systems causes that the HRT is not long enough to carry out the hydrolysis of retained particulates when temperature is lower than 20 C. Therefore, solids accumulate in the granular sludge bed which deteriorates the overall methanogenic activity and the reactor performance. Several strategies could be applied to avoid this deterioration of sludge quality: a) Wastewater may be pretreated to remove suspended solids (Aiyuk et al., 2006); b) the use of a two step system (hydrolytic reactor + UASB reactor) (Álvarez et al., 2008); c) The use of anaerobic membrane bioreactors (Liao et al., 2006). Since anaerobic digestion is only able to remove organic matter, the CANON process should be suitable to remove ammonia (Figure 16). This process was only applied to effluents with temperatures ranged between 30-40 ºC since it is the optimum range of temperature for Anammox bacteria. However, some works showed that the Anammox process could be successfully operated at temperatures around 20 ºC (Dosta et al., 2008) and nitrogen removal rates up to 1.1 kg NH4+-N /(m3·d) can be achieved at this temperature (Vázquez-Padín et al., 2009a; 2009b).

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This proposed scheme would allow: a) Reducing the size of the WWTP; b) Obtaining a positive net balance of energy in the WWTP; c) Reducing both sludge generation and CO2 emissions.

4.3. Ground Water Bioremediation In the last years, nitrate levels in ground waters exceeding the European Regulation (11.3 mg NO3--N/L) were observed. The conventional method to remove nitrate is ionic exchange although the application at full scale of reverse osmosis also gave good results. Nevertheless, both processes generate a residual stream which needs a postreatment. An alternative to these technologies is the denitrification. In the case of heterotrophic denitrification, organic matter (ethanol or methanol) must be added as electron donor that leads to a secondary contamination. This can be avoided if nitrate removal is done by autotrophic bacteria using elemental sulfur since it is not a toxic compound and it is insoluble in water. This process will generate sulfate and is recommended to apply to ground water with low endogenous sulfate levels to avoid sulfate concentrations higher than 400 mg SO4-2/L. The application of autotrophic denitrification to ground water has been limited by the low biomass retention. Therefore, recent works are focused on combining this process with membrane (McAdam y Judd, 2006) or biofilm technologies (Soares, 2002) to achieve a complete retention of the biomass. The configurations proposed are the following (Figure 17): (a) Bioreactor with extractive membrane: In this configuration, nitrate is extracted from water by molecular diffusion through the membrane to a stream containing both denitrifying biomass and electron donor (Figure 17a). Biogas FeCl3 + Flocculant (SS and P removal)

Influent Effluent

Primary settler

Sludge

UASB reactor

CANON system

Figure 16. Municipal wastewater treatment using a CANON system to remove nitrogen.

(b) Bioreactor with filtration membrane: Denitrifying biomass is mixed with polluted ground water and electron donor. In this case, the membrane is used to separate biomass from treated water by application of pressure (Figure 17b). (c) Biofilm reactor: Elemental sulfur particles could be used as both electron donor and support of autotrophic denitrifying biomass. A column filled with elemental sulfur granules and operated in an upflow mode could be a system very simple, stable and easy to maintain (Figure 17c).

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A

Treated water

Denitrifying bacteria Sulfur

Water with nitrate

Membrane Denitrifying biomass Residual water

B Water with nitrate

Denitrifying biomass

Water with nitrate

Treated water

Membrane

Denitrifying bacteria Sulfur Water with nitrate

Treated water

Membrane

Residual water

Membrane

C

Denitrifying bacteria

Water with sulfate

Water with nitrate

Sulfur

Water with nitrate

Figure 17. Systems to remove nitrate from ground water: a) bioreactor with extractive membrane, b) bioreactor with filtration membrane and c) biofilm reactor.

4.4. Nitrate Removal from Recirculating Aquaculture Systems Factors such as limitations of water quality, land costs, disposal requirements and environmental impact are driving the aquaculture sector to more intensive practices. The use of recirculating systems allows reducing water used and disposed during aquaculture activities. Besides, it has another advantages: a) Saving of pumping costs; b) Control of pH and temperature which optimize fish production; c) Presence of pathogens is minimized which reduces mortality during the broodstock stage. Since ammonia is toxic for fish at concentrations higher than 1.5 mg NH4+-N/L, this compound must be removed by a nitrifying biofilter to avoid its accumulation in the system. Ammonia is oxidized into nitrate which is less toxic for fishes, its recommended limit being around 50 mg NO3--N/L. However its effect depends on the specie and growth stage and, therefore, its removal is advisable. The use of denitrifying biofilter with elemental sulfur would be the most suitable option to maintain nitrate concentration as low as possible (Figure 18). The sulfate generated during the autotrophic denitrification would cause neither environmental nor toxicity problems when marine species are cultured (Vidal et al., 2002).

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5. CONCLUSIONS Implantation of bioaugmentation, partial nitrification or anammox processes in the reject water stream of WWTPs supposes an economical and feasible alternative to improve effluent quality in terms of nitrogen content. The most suitable technology must be chosen depending on the WWTP operational conditions. Perspectives of advanced nitrogen removal processes application are very promising in fields such as wastewater, drinking water or aquaculture systems. Fresh water

Sedimentation zone

Fish culture tank

UV unit Effluent

Solid wastes

Recycling water

NH4+ Organic matter

Denitrifying biofilter SO4-2

NO3-

Nitrifying biofilter

Figure 18. Recirculating aquaculture system.

ACKNOWLEDGMENTS This work was funded by the Spanish Government (TOGRANSYS project CTQ200806792-C02-01/PPQ and NOVEDAR_Consolider project CSD2007-00055).

6. REFERENCES Abma, W., Schultz, C., Mulder, J. M., van Loosdrecht, M., van der Star, W., Strous, M. & Tokutomi, T. (2007). The advance of Anammox. Water 21, Feb. 2007, 36-37. Aiyuk, S., Forrez, I., Lieven, K., van Andel, A. & Verstraete, W. (2006). Anaerobic and complementary treatment of domestic sewage in regions with hot climates- A review. Bioresource Technology, 97, 2225-2241. Álvarez, J. A., Armstrong, E., Gómez, M. & Soto, M. (2008). Anaerobic treatment of lowstrength municipal wastewater by a two-stage pilot plant under psychrophilic conditions. Bioresource Technology, 99, 7051-7062.

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Berends, D. H. J. G., Salem, S., van der Roest, H. F. & van Loodsdrecht, M. C. M. (2005). Boosting nitrification with BABE technology return sludge. Water Science and Technology, 52(4), 63-70. Campos, J. L., Carvalho, S., Portela, R., Mosquera-Corral, A. & Méndez, R. (2008). Kinetics of denitrification using sulfur compounds: Effects of S/N ratio, endogenous and exogenous compounds. Bioresource Technology, 99, 1293-1299. Dapena-Mora, A., Campos, J. L., Mosquera-Corral, A., Jetten, M. S. M. & Méndez, R. (2004). Stability of the Anammox process in a gas-lift reactor and a SBR. Journal of Biotechnology, 110, 159-170. Dapena-Mora, A., Fernández, I., Campos, J. L., Mosquera-Corral, A., Méndez, R. & Jetten, M. S. M. (2007). Evaluation of activity and inhibition effects on Anammox process by batch tests based on the nitrogen gas production. Enzyme and Microbial Technology, 40, 859-865. Dosta, J., Fernández, I., Vázquez-Padín, J. R., Mosquera-Corral, A., Campos, J. L., Mata, J. & Méndez, R. (2008). Short- and long-term effects of temperature on the Anammox process. Journal of Hazardous Materials, 154, 688-693. Fajardo, C., Mosquera-Corral, A., Campos, J. L. & Méndez, R. (2008). Depuración conjunta de aguas ricas en nitratos y efluentes con compuestos reducidos del azufre. Retema, 127, 38-51. Fux, C. & Siegrist, H. (2004). Nitrogen removal from sludge digester liquids by nitrification/denitrification or partial nitritation/anammox: environmental and economical considerations. Water Science and Technology, 50(10), 19-26. Garuti, G., Giordano, A. & Pirozzoli, F. (2001). Full-scale ANANOX system performance. Water SA, 27, 189-197. Gijzen, H. J. (2001). Anaerobes, aerobes and phototrophs - A winning team for wastewater management. Water Science and Technology, 44, 123-132. Gong, Z., Yang, F. L., Liu, S. T., Bao, H., Hu, S. W. & Furukawa, K. J. (2007). Feasibility of a membrane aerated biofilm reactor to achieve single-stage autotrophic nitrogen removal based on Anammox. Chemosphere, 69, 776-784. Hellinga, C., Schellen, A. A. J. C., Mulder, J. W., Van Loosdrecht, M. C. M. & Heijnen, J. J. (1998). The Sharon process: An innovative method for nitrogen removal from ammonium-rich waste water. Water Science and Technology, 37(9), 135-142. Kampschreur, M. J., van der Star, W. R. L., Wielders, H. A., Mulder, J. W., Jetten, M. S. M. & van Loosdrecht, M. C. M. (2008). Dynamics of nitric oxide and nitrous oxide emission during full-scale reject water treatment. Water Research, 42, 812-826. Khin, T. & Annachhatre, P. (2004). Novel microbial nitrogen removal processes. Biotechnology Advances, 22, 519-532. Kroiss, H. (2004). What is the potential for utilizing the resources in sludge? Water Science and Technology, 49(10), 1-10. Liao, B. Q., Kraemer, J. T. & Badgley, D. M. (2006). Anaerobic Membrane Bioreactors: Applications and Research Directions. Critical Reviews in Environmental Science and Technology, 36, 489-530. Liu, S., Yang, F., Gong, Z. & Su, Z. (2008). Assessment of the positive effect of salinity on the nitrogen removal performance and microbial composition during the start-up of CANON process. Applied Microbiology and Biotechnology, 80, 339-348.

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McAdam, E. J. & Judd, S. J. (2006). A review of membrane biorreactor potencial for nitrate removal from drinking water. Desalination, 196, 135-148. Mosquera-Corral, A., González, F., Campos, J. L. & Méndez, R. (2005). Partial nitrification in a SHARON reactor in the presence of salts and organic carbon compounds. Process Biochemistry, 40, 3109-3118. Park, S. & Bae, W. (2009). Modeling kinetics of ammonium oxidation and nitrite oxidation under simultaneous inhibition by free ammonia and free nitrous acid. Process Biochemistry, 44, 631-640. Pynaert, K., Smets, B. F., Beheydt, D. & Verstraete, W. (2004). Start-up of autotrophic nitrogen removal reactors via sequential biocatalyst addition. Environmental Science and Technology, 38, 1228-1235. Salem, S., Berends, D. H. J. G., Heijnen, J. J. & Van Loodsdrecht, M. C. M. (2002a). Bioaugmentation by nitrification with return sludge. Water Research, 37, 1794-1804. Salem, S., Berends, D., Heijnen, J. J. & van Loosdrecht, M. C. M. (2002b). Model-based evaluation of a new upgrading concept for N-removal. Water Science and Technology, 45(6), 169-176. Salem, S., Berends, D. H. J. G., van der Roest, H. F., van der Kuij, R. J. & Van Loodsdrecht, M. C. M. (2004). Full scale application of the BABE technology. Water Science and Technology, 50(7), 87-96. Siegrist, H., Salzgeber, D., Eugster, J. & Joss, A. (2008). Anammox brings WWTP closer to energy autarky due to increased biogas production and reduced aeration energy for Nremoval. Water Science and Technology, 57(3), 383-388. Sliekers, A. O., Tirad, K. A., Abma, W., Kuenen, J. G. & Jetten, M. S. M. (2003). CANON and Anammox in a gas-lift reactor. FEMS Microbiology Letters, 218, 339-344. Sliekers, O., Derwort, N., Campos-Gomez, J. L., Strous, M., Kuenen, J. G. & Jetten, M. S. M. (2002). Completely autotrophic nitrogen removal over nitrite in a single reactor. Water Research, 36, 2475-2482. Soares, M. I. M. (2002). Denitrification of groundwater with elemental sulphur. Water Research, 36, 1392-1395. Tanaka, K., Sumino, T., Nakamura, H., Ogasawara, T. & Emori, H. (1996). Application of nitrification by cells immobilized in polyethylene glycol. In: R. H. Wijjfels, R. M. Buittelaar, C. Bucke, & J. Tramper (Eds.), Immobilized Cells: Basics and Applications. Amsterdam: Elsevier Science, 622-632. Van der Star, W. R. L., Abma, W. R., Blommers, D., Mulder, J. W., Tokutomi, T., Strous, M., Picioreanu, C. & van Loosdrecht, M. C. M. (2007). Startup of reactors for anoxic ammonium oxidation: Experiences from the first full-scale anammox reactor in Rotterdam. Water Research, 41, 4149-4163. Van Dongen, U., Jetten, M. S. M. & Loosdrecht, M. C. M. (2001). The SHARON®Anammox® process for treatment of ammonium rich wastewater. Water Science and Technology, 44(1), 153-160. Van Kempen, R., Mulder, J. W., Uijterlinde, C. A. & van Loosdrecht, M. C. M. (2001). Overview: full scale experience of the SHARON process for treatment of rejection water of digested sludge dewatering. Water Science and Technology, 44(1), 145-52. Van Loosdrecht, M. C. M. & Salem, S. (2006). Biological treatment of sludge digester liquids. Water Science and Technolology, 53(12), 11-20.

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Vázquez-Padín, J. R., Fernández, I., Figueroa, M., Mosquera-Corral, A., Campos, J. L. & Méndez, R. (2009a). Applications of Anammox based processes to treat anaerobic digester supernatant at room temperature. Bioresource Technology, 100, 2988-2994. Vázquez-Padín, J. R., Pozo, M. J., Jarpa, M., Figueroa, M., Franco, A., Mosquera-Corral, A., Campos, J. L. & Méndez, R. (2009b). Treatment of anaerobic sludge digester effluents by the CANON process in an air pulsing SBR. Journal of Hazardous Materials, 166, 336-341. Vidal, S., Rocha, C. & Galvao, H. (2002). A comparison of organic and inorganic carbon controls over biological denitrification in aquaria. Chemosphere, 48, 445-451. Wett, B. (2006). Solved upscaling problems for implementing deammonification of rejection water. Water Science and Technology, 53(12), 121-128. Wett, B. (2007). Development and implementation of a robust deammonification process. Water Science and Technology, 56(7), 81-88. Wiesmann (1994). Biological Nitrogen Removal from Wastewater. In: Fletcher A. (ed.), Advances in Biochemical Engineering Biotechnology, vol. 51. Spinger-Verlag, Berlín. 113-154. Windey, K., de Bo, I. & Verstraete, W. (2005). Oxygen-limited autotrophic nitrificationdenitrification (OLAND) in a rotating biological contactor treating high-salinity wastewater. Water Research, 39, 4512-4520.

In: Fluid Waste Disposal Editor: Kay W. Canton, pp. 179-200

ISBN: 978-1-60741-915-0 © 2010 Nova Science Publishers, Inc.

Chapter 7

APPLICATION OF MICROBIAL MELANOIDINDECOMPOSING ACTIVITY (MDA) FOR TREATMENT OF MOLASSES WASTEWATER Suntud Sirianuntapiboon* and Sadahiro Ohmomo Department of Environmental Technology, School of Energy, Environment and Materials, King Mongkut‘s University of Technology, Thonburi (KUMTT), Bangmod, Thung Khru, Bangkok 10140, Thailand.

ABSTRACT This review will discuss the melanoidin-decomposing activity (MDA) among microorganisms. The focus will be on the potential use of the microbial-MDA to treat the wastewater discharged from factories using molasses as the raw material (molasses wastewater: MWW) because molasses is one of the most useful raw materials in various types of industries, such as the fermentation and animal feed industries. However, the wastewater discharged from factories using molasses contains a large amount of dark brown pigment, melanoidin pigment: MP, which is poorly decomposed and/or decolorized by normal biological treatment processes, such as the activated sludge or anaerobic treatment systems (anaerobic pond or anaerobic contact digester), because, the microorganisms in those wastewater treatment systems showed very poor MDA. The distribution of MDA among microorganisms and the mechanism of decomposing activities, in particular, were reviewed. Also, the application of the isolated strains having the MDA to treat molasses wastewater in the wastewater treatment plant was tested.

Keywords: Melanoidin, Molasses, Molasses wastewater, Decolorization, Alcohol distillation, Microbial decolorization, Chemical decolorization.

*

Corresponding author: E-mail: [email protected]

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INTRODUCTION A large body of scientific research has been conducted on the conversion of renewable biomass to useful materials for political and practical concerns over the state of the environment (Underfolker and Hickely, 1954). In particular, bio-fuel from renewable biomass is expected to help alleviate the ongoing energy crisis. Molasses, a by-product from sugar cane or sugar beat in sugar industries, is one of the largest sources of biomass and a very important material worldwide. Molasses consists of about 50% sugar (reducing sugar) or related substrates, about 10% non-sugar organic substances, and about 10% minerals (Underfolker and Hickely, 1954). Due to these components, molasses was widely used in various fermentation industries, such as alcohol fermentation, amino acid fermentation, antibiotics fermentation, and baker‘s yeast fermentation, (Chang and Yang, 1973; Chaung and Lai, 1978), as a low-cost and readily available raw material when diluted with water. Molasses from sugar cane is mainly produced in tropical areas of the world, especially in south-east Asia (Philippines, Indonesia, Thailand, etc). Wastewater from the fermentation processes using molasses is densely colored by molasses pigment, called melanoidin pigment (MP), and contains a large amount of organic matters, which leads to high biological oxygen demand (BOD5) and chemical oxygen demand (COD) values (Sirianuntapiboon et al., 1988a; Antonia et al, 2000). Therefore, the wastewater can be treated by normal biological treatment processes, such as the activated sludge system, aerated lagoon, or anaerobic pond, to remove the organic matter. However, MP is poorly decomposed and still remains in the wastewater after treatment by above processes. No suitable method for the treatment of large amounts of this type of wastewater has been developed yet, so this is a problem that still needs to be solved. For example, this problem has led to an increase in the production cost of the ethyl alcohol fermentation process from molasses, because about 75% of potential energy in ethyl alcohol is wasted due to the popular treatment process as concentration and combustion of the wastewater (Chaung and Lai, 1978; Chang and Yang, 1973). Therefore, the development of a low-cost and simple wastewater treatment system that utilizes microbes to decompose and decolorize MP is urgently needed. In this paper, the MP-decomposing activities (MDA) in microorganisms are reviewed with a focus on the distribution of MDA in microorganisms, the mechanisms of their activities, and the application process for the treatment of wastewater from the factories using molasses.

1. MELANOIDIN PIGMENTS (MP) (Monica et al., 2004; Kwak et al, 2005; Kato and Hayase, 2002; Fogliano et al, 1999; Yaylayan et al, 1998): MP is synthesized from carbonyl compounds, such as sugar and amino compounds, amino acids, or proteins (Monica et al, 2004; Kato and Hayase, 2002), as shown in Figure 1. This is a non-enzymatic browning reaction and a type of amino-carbonyl reaction called Maillard reaction (Monica et al., 2004). This reaction is promoted under the alkaline condition and the coloration of MP changes from yellow to dark brown with increased MPcondensation (Fogliano et al, 1999). MP grows large molecular weight substances (more than

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100,000 Dalton) and the condensed-MP precipitates under acidic conditions of pH less than 3. The color-density of MP solution under acidic conditions (pH 5.0) is weakened by about 20% in comparison with the alkaline condition and shows a maximum absorption at a wavelength of 475 nm (Ohmomo et al, 1985a; Sirianuntapiboon et al, 1988a). The colored substances in some foods, such as Shoyu and Miso (Japanese seasoning), as well as molasses, are typical MPs (Ohmomo et al., 1985a).

2. SCREENING OF MICROORGANISMS HAVING MP-REMOVAL ACTIVITY (MPRA) (Ohmomo et al., 1988a; Watanabe et al., 1982; Sirianuntapiboon et al., 1988a): MP is known as a poorly biologically-decomposed substance because of its complicated-structure (Monica et al., 2004; Kato and Hayase, 2002 ). Several researchers tried to isolate the microorganisms which have MPRA as shown in Table 1. It was found that several types of microorganisms such as mushroom (Watanabe, et al., 1982; Ohmomo, et al., 1985a; Kumar, et al., 1998; Miyata, et al., 2000), mold (Miyata, et al., 2000; Ohmomo et al, 1987a; Sirianuntapiboon et al, 1988b; Fahy et al, 1997; Kim and Shoda, 1999; Dahiya et al, 2001a; Jimenez et al, 2003), yeast (Sirianuntapiboon et al, 2004b) and bacteria (Ohmomo et al, 1988b; Francisca et al, 2001; Sirianuntapiboon et al, 2004a; Kumar and Chamdra, 2006) showed the MPRA. Some of them showed very strong MP-adsorption ability (Ohmomo et al, 1988b; Watanabe et al, 1982) or MP-degradation ability (Fahy et al, 1997) while the other showed both MP-adsorption and MP-degradation abilities (Sirianuntapiboon et al, 1995). It was understood that microorganisms having MDA were not common habitants. At first, the microbial-MDA was screened among the white-rod fungi (mushroom) due to their lignindecomposing activity by a research group of Kyushu University, Japan (Watanabe et al., 1982). They succeeded in isolating a strain Coriolus sp. No. 20 and suggested that the MDA in this strain was led by a sorbose oxidizing enzyme: sorbose oxidase (Watanabe et al., 1982). This was the first report for microbial MDA in the world. After this, our research group also found the MDA in some white-rod fungi to be especially strong in Coriolus versicolor Ps4a (Ohmomo et al., 1985a; Ohmomo et al., 1985b; Ohmomo et al., 1985c; Aoshima et al., 1985), and succeeded in confirming that there was MDA among various groups of microorganisms, such as fungi, bacteria, and yeast (Ohmomo et al., 1987a; Ohmomo et al., 1987b; Ohmomo et al., 1987c; Sirianuntapiboon, et al., 1988a; Sirianuntapiboon, et al., 1988b; Sirianuntapiboon, et al., 2004a; Sirianuntapiboon, et al., 2004b).

2.1. MP and MP Solution Preparation Two kinds of MP solutions were used in screening the microorganisms for MDA and MDA determination as natural-MP (NMP) and synthetic-MP (SMP) solutions. The NMP solution was fundamentally prepared from the MWW (Ohmomo et al., 1987a; Sirianuntapiboon et al., 1988a). Two kinds of MWW as stillage from an alcohol factory (U-MWW) and treated-MWW especially from an anaerobic pond (An-MWW) could be used for preparation of the NMP solution. However, the molecular weight distribution of NMP solution was not always

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equalized, because of the conditions of the sugar making process, fermentation of molasses, and treatment process and condition of MWW (Sirianuntapiboon and Chairattanawan, 1998). Due to these uncertainties, synthetic-MP (SMP) was more widely used for the screening of microbial-MDA. SMP was synthesized by heating the solution containing 1 mol/L glucose, 1 mol/L glycine and 0.5 mol/L Na2CO3 at 1210C for 3 hr (Sirianuntapiboon, et al., 1988a; Ohmomo et al., 1985a). After heating, the solution was adjusted to a pH of 7.0 with 1.0 mol/L NaOH solution and ultra-filtrated using membrane filters of molecular weight cut-offs between 1,000 Dalton and 10,000 Dalton. The fractions having molecular weights from 1,000 to 10,000 Dalton was harvested and freeze-dried to make a SMP powder. The solution of SMP is prepared as giving an optical density of 3.5 (OD=3.5) at a wavelength 475 nm in 0.1 mol/L acetate buffer (pH 5.0) before being used in the experiments.

2.2. Screening Methods for Isolation of Microbial Strain Having MDA The medium containing SMP or NMP was used for screening the microorganisms having MDA. Fungal and bacterial strains were isolated by using media containing MP. The testedmicroorganisms were inoculated on the surface of an agar medium suitable for growth (the media contained MP) and were cultured to make a colony. If the tested-microorganism had MDA, a clear zone around the colony was formed (Sirianuntapiboon, et al., 1988a). Furthermore, the microorganisms forming the clear zone around the colony was cultured in a liquid medium suitable for the growth and the medium color intensities before and after cultivation were compared in order to calculate the decolorization yield (Sirianuntapiboon, et al., 1988a). If the tested-strain produced organic acids and reduced medium pH, the color density of the culture filtrate was reduced due to the acidic-pH. Therefore, the color intensity of culture filtrate should be measured after dilution with 0.1 mol/L acetate buffer (pH 5.0) to prevent the error of the color intensity reduction in acidic-pH condition (Sirianuntapiboon, et al., 1988a).

Figure 1. Outline for the formation of MP: Maillard‘s reaction (Monica et al., 2004)

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Table 1. MP removal mechanisms of various types of microorganisms. Type of Microorganisms Mushroom

Name of microorganism (Genus and species) Coriolus sp. No.20 Coriolus versicolor Ps4a

Coriolus versicolor

Coriolus hirsutus

Mold

Aspergillus fumigatus G-2-6

Rhizoctonia sp. D90

Aspergillus oryzae Phanerochaete chrysosporium

Geotrichum candidum Dec 1 Phanerochaete chrysosporium JAG-40 Penicillium decumbens Flavodon flavus

Yeast

Citeromyces sp. WR-43-6

Bacteria

Lactobacillus hirgardii

Oscillatoria boryana BDU 92181 Acetogenic bacteria BP103

MP removal mechanisms Decolorization by sorbose oxidase enzyme Decolorization by intracellular enzyme (inducible enzyme by MP) Decolorization by sorbose oxidase, sugar oxidase and manganese dependent peroxidase MP degradation by Intracellular enzyme (sugar oxidase) MP degradation by Intracellular enzyme (Inducible enzyme by MP) MP removal by adsorption, absorption and decomposition MP removal by adsorption MP decomposing enzyme (showed highest ability at the stationary phase of growth curve) MP degradation by peroxidase enzyme MP degradation by extracellular enzyme MP removal by degradation and/or adsorption MP degradation by glucose oxidase and hydrogen peroxide enzymes MP degradation by sugar oxidase enzyme. MP removal by assimilation and degradation peroxidase enzyme MP removal by assimilation and peroxidase enzyme MP degradation by sugar oxidase enzyme

References Watanabe et al., 1982 Ohmomo et al., 1985

Kumar et al., 1998

Miyata et al., 2000

Ohmomo et.al., 1987

Sirianuntapiboon et al., 1988 Fahy et al., 1997; Ohmomo et al, Kim and Shoda, 1999

Dahiya et al., 2001 Jimenez et al., 2003 Raghukumar et al., 2004 Raghumar and Mohandass, 2001 Sirianuntapiboon et al., 2004 Ohmomo et al., 1988

Francisca et al., 2001 Sirianuntapiboon et al.,2004

2.3. MDA in the Fungal Strain Several fungal strains belonging to the classes of Basidiomycetes, Ascomycetes and Dueteromycetes showed MPRA were isolated. The strain belonging to class Basidiomycetes was first isolated as the MDA strain. Most of the white-rod fungi strains that have the ability to decompose lignin, showed MDA, and especially strong MDA were detected in Coriolus versicolor Ps4a, Coriolus hirsutus, Fomitopsis cytisina, Irpex lactens, and Pleurotus ostreatus

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(Watanabe, et al., 1982; Ohmomo, et al., 1985a; Raghukumar and Rivonkar, 2001; Miyata, et al., 1998; Fitzgibbon, 1998). But, the strains of brown-rod fungi having the ability to decompose cellulose never showed the MDA (Aoshima et al., 1985). Most of the fungi having MDA grew well on shaking cultures using a medium containing glucose, sucrose, or maltose and showed very strong MDA (Watanabe, et al., 1982). However, the MDA was weak on the cultures using xylose or arabinose as the carbon source, while the growth rate was high. Additionally, nitrogen sources were also affected to growth and MDA. Organic nitrogen sources, such as peptone and casamino acid were the best nitrogen source in obtaining a high growth rate and strong MDA. Ammonium salt was also good for growth, but it gave only half level of MDA of that with peptone. Nitrate salts gave poor growth and weak MDA. The highest decolorization yield (75-80%) of Coriolus versicolor Ps4a was obtained in a shaking culture using a MP-medium containing 5% glucose and 0.5% peptone at 30oC for 4-6 days, and the MDA was the decomposition of MP by decreasing the molecular weight of the MP (Ohmomo et al., 1985b). A research group at Kobe University also screened for MDA among white-rod fungi and detected a strong MDA from Coriolus pubescens, Hirschioporus fuscoviolaceus, Polyporellus brumalis, etc., and at the same time, they found strong browning activity in some unidentified fungi (Tamaki et al., 1985). Moreover, Trametes versicolor (Benito et al., 1997), Phanerochaetes chrysosporium (Fahy et al., 1997; Kumer et al., 1998; Kumer et al, 1997; Fitgibbon et al., 1998; Dahiya, 2001a), Coriolus hirsutus (Miyata et al., 1998), Flavodon flavus (Raghkumar et al., 2004) and others have since been reported as having MDA. For the screening of thermophilic fungi, the stain of Ascomycetes, mainly belonging to the genus Aspergillus, strain G-2-6 was isolated as showing the strongest MDA and identified as Aspergillus fumigatus. The strain gave the maximum decolorization yield of 75% on a shaking culture using a medium containing glycerin and peptone as the carbon and nitrogen sources, respectively, at 45oC for 3 days, and the MP-removal mechanism was the decomposition of MP (Ohmomo et al., 1987a). At the same time, strain Y-2-32 was also isolated as showing the strongest MDA and identified as Aspergillus oryzae. However, the MDA of this strain was not the MP-decomposition mechanism, but MP was only adsorbed onto the surface of mycelia (Ohmomo et al., 1988). The ability to adsorb MP among living mycelia and dead mycelia was almost the same and was recovered after washing of mycelia with buffer solution. The reuse of mycelia after washing was possible (Ohmomo et al., 1988b). Aspergillus niger 180 was also selected (Miranda et al., 1996) and the MDA of Aspergillus niger when combined with Penicillium decumbens and Penicillium lignorum was also confirmed (Antonia et al., 2003). Additionally, Rhizoctonia sp. D90 (=Mycelia sterilia D90), which belongs to class Deuteromycetes, was screened as having strong MDA (Sirianuntapiboon et al., 1988a, Sirianuntapiboon et al., 1988b; Sirianuntapiboon et al., 1995). This strain gave the maximum decolorization and COD removal yields of more than 90% and 80%, respectively (Sirianuntapiboon et al., 1988a; Sirianuntapiboon et al., 1988b; Sirianuntapiboon, 1995). The MDA of Geotrichum candidum (Kim and shoda, 1999), Oscillatoria boryana (Kalavathi et al., 2001) and Paecilomyces canadensis (Terasawa et al., 2000) were also detected.

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2.4. MDA in the Bacterial Strain Both aerobic bacteria and anaerobic bacteria strains having MDA were isolated (Ohmomo et al, 1987b; Ohmomo et al, 1988a; Sirianuntapiboon et al, 2004b), and both strains were applied to the conventional wastewater treatment systems (in the laboratory scale), activated sludge system and anaerobic treatment system, respectively (Kumar et al, 1997; Mohana et al, 2007; Ghosh et al, 2002). Lactic acid bacteria strains were isolated in order to find a strain of anaerobic bacteria with MDA that could be applied in an anaerobic treatment system (Kumar et al., 1997; Kumar et al., 1998; Mohana et al, 2007; Ohmomo et al, 1987b). Among them, a hetero-fermentative strain W-NS identified as Lactobacillus hilgardii showed the strongest MDA and the maximum decolorization yield was about 30% for sugar cane molasses, about 40% for beat molasses, and about 65% for synthetic glycine-glucoseMP under the presence of 1% glucose at 35-40oC (Ohmomo et al., 1987b). Furthermore, acetogenic bacteria BP103, an aerobic bacteria, was isolated from Thailand showed a strong MDA. This strain decolorized 75-80% of molasses wastewater under the presence of 3% glucose and 0.5% peptone at 30oC (Sirianuntapiboon et al., 2004a). The MDA was also detected in Bacillus smithii, which decolorized about 36% of molasses wastewater at 55oC within 20 days (Nakajima-Kambe et al., 1999). The MDA were also detected in Pseudomonas fluorescens (Jagroop et al., 2001; Dahiya et al., 2001), Pseudomonas putida (Ghosh et al., 2002) and some strains belonging to the genus Methanothrix and Methanosarcina (Boopathy and Tilche, 1991; Boopathy, 1992). In addition, the mixed culture of Streptomyces warraensis and Basidiomycetes fungi was also tested for decolorization of MWW (Terasawa et al., 2000).

2.5. MDA in the Yeast Strain Our research group tried to isolate the yeast strain having MDA from several sources of fruit and soil in Thailand. Citeromyces sp. WR-43-6 was first isolated as the MDA strain in the yeast group. This strain gave a maximum decolorization yield of more than 70% and at the same time, BOD5 and COD were removed by more than 76% and 98%, respectively (Sirianuntapiboon et al., 2004b).

3. MECHANISM OF MICROBIAL MDA The microbial-MDA or removal mechanisms were investigated by several research groups. The fungal enzyme system related to the MDA was partially purified in Coriolus sp. No.20 (Watanabe et al., 1982) and Coriolus versicolor Ps4a (Ohmomo et al., 1988a). Some steps, such as adsorption of MP onto the cell surface and incorporation of MP in the cell were suggested in fungi (Sirianuntapiboon et al, 1998; Ohmomo et al, 1988b). However, the mechanism of microbial MDA was hardly elucidated. Because, some isolated-strains showed strongly on the MP-degradation mechanism, while the others showed strongly on the MPadsorption mechanism. For the observation of MDA mechanism of the Rhizoctonia sp. D-90 strain, MDA mechanism by Rhizoctonia sp. D-90 involved in absorption of the MP by the

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cells as a macromolecule and its intracellular accumulation in the cytoplasm and around the cell membrane as a MP complex, which was then gradually degraded by an intracellular enzyme. However, the main-MP decolorization or removal mechanisms of each MPdecolorization strain was different. The details of MDA and MP-removal mechanisms are described below:

3.1. Microbial-Decomposition of MP Few papers related to the mechanism of the enzyme system in MDA have been published, despite the numerous publications on the decolorization and decomposition of MP from molasses. As a first step in resolving the mechanisms in microbial MDA, sorbose oxidase was partially purified from the mycelia of white-rod-fungi, Coriolus sp. No. 20 (Watanabe et al., 1982). It was thought that the mechanism of MDA by this enzyme system would act by oxidizing sorbose and release activated-oxygen (oxygen radical) to decompose MP. However, no clear evidence for this system was found. Two enzymes, P-III and P-IV, related to the decolorization of MP were partially purified from the mycelia of Coriolus versicolor Ps4a (Ohmomo et al., 1985a; Ohmomo et al, 1985b). Enzyme P-III with a molecular weight of about 50,000 Dalton gave the decolorization yield of about 11% under aerobic conditions and in the presence of glucose. However, enzyme P-IV with a molecular weight of about 45,000 Dalton gave the decolorization yield of about 13% under anoxic conditions and without glucose. The maximum decolorization yield of each enzyme was low; however, the multiplicative effect (Figure 2) for decolorization with both enzymes was observed with the decolorization yield of about 40%, which was higher than that calculated sum of decolorization yields of both enzymes of 24%. Decolorization activity of enzyme PIII and PIV was 11% and 13%, respectively. Furthermore, lactic acid and various amino acids were detected as the reaction products of these enzymes (Ohmomo et al., 1985c). Conversely, the relation of manganese-dependent peroxidase to the decomposition of MP was suggested in Coriolus hirsutus (Miyata et al., 1998, 2000) and Flavodon flavus (Raghkumar and Rivonkar, 2001, Raghukumar et al, 2004). For the determination of molecular weight distribution of MP decomposed by microorganisms, it was found the molecular weight of MP of the treated SMP or NMP solutions were shifted to smaller molecular weight fractions than that of the initial solutions, as shown in Figure 3, and this shift was also detected in the case of fungi and bacteria (Ohmomo et al., 1988; Sirianuntapiboon et al., 1988b). In addition, the production of hydrogen peroxide and activated oxygen by photo-synthetic cyanobacteria closely participated in the decomposition of MP by Oscillatoria boryana BDU 92181 (Kalavathi et al., 2001). Further, the induction at low level MP (10 g/liter) and the inhibition at high level MP (20 g/liter) for the production of peroxidase was reported in Geotricum candidum (Lee et al., 2000).

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A: shows the decolorization by P-IV after complete decolorization by P-III. B: shows the decolorization by P-III after complete decolorization by P-IV. C: shows the decolorization by a mixture of P-III and P-IV. Symbols: , actual decolorization yield; , decolorization yield (calculated from the individual decolorization yields of P-III and P-IV); ↓, shows the addition of an enzyme. Figure 2. Multiplicative effect between enzyme P-III and P-IV for the MP Decolorization (Ohmomo et al., 1985a).

Chromatograms of MP solution were obtained by using gel filtration on a Sephadex G-50 column. Symbols: , initial MP solution; , solution treated by Citeromyces sp. WR-43-6. Figure 3. Molecular weight distribution in MP solution before and after decolorization by Citeromyces sp. WR-43-6 (Sirianuntapiboon et al., 2004b).

3.2. Microbial-Adsorption of MP The adsorption of MP onto the cell surface was suggested in Aspergillus oryzae Y-2-32, which strongly decolorized MWW. The decolorization of this strain was due to the adsorption of MP onto the cell surface, and its yield depended on the amounts of cell mass. However, the MP adsorption ability disappeared when washed with 0.1% Tween 80 solution or 0.1% SDS

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solution. This means that there is a relation between the cell surface components, such as muco-polysaccharides, and MP adsorption (Ohmomo et al.,a 1988b). The adsorption of MP onto the cell surface should be the first step of the MDA, and this strain has no next step, such as incorporation of MP into the cell and/or decomposition of MP. However, the electron microscopic observation for the adsorption and incorporation of MP onto the cell (cell membrane and cytoplasm) of Rhizoctonia sp. D90 (= Mycelia sterilia D90) was reported (Sirianuntapiboon et al., 1995). Tremetes versicolor gave a strong decolorization yield, maximum 80%, and about 10% of the yield was due to the adsorption onto the cell surface (Benito et al., 1997). These reports could suggest that the MDA displays a two-step reaction of adsorption of MP onto the cell surface and incorporation of MP into the cell as shown in Figure 4 and Figure 5.

a

b

a

a: Cross-section of 7-day-old mycelium, grown in synthetic melanoidin medium, showing electrondense materials distributed in the cytoplasm. b: Cross-section of 7-day-old mycelium, grown in potato dextrose medium, showing well-defined cell organelles such b as the cell wall (cw) and the cell membrane (cm). Figure 4. Electron Micrographs of Rhizoctonia Sp. D-90 in SMP medium and potato dextrose medium (Sirianuntapiboon et al., 1995).

a

b

a: Cross-section of mycelium that had been grown in potato dextrose medium for 7 days (100,000 x magnification), showing the clear cytoplasm and cell membrane (cm). b: Cross-section of mycelium that had been grown in potato dextrose medium for 7 days and then in SMP-medium for another 4 days (100,000 x magnification), showing electron-dense materials distributed in b the cytoplasm. Figure 5. Electron micrographs of Rhizoctonia sp. D-90 collected at various stages of cultivation (Sirianuntapiboon et al., 1995).

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Continuous decolorization was carried out in a bubbling column (diameter 26 mm x length 400 mm) at 30 0C with aeration. The column contained 100 ml of waste water and 25 g of immobilized mycelia (corresponding to 0.8 g dry mycelial weight). → shows the start of continuous decolorization at a dilution rate of 0.022 hr-1. Figure 6. Continuous decolorization by immobilized mycelia of Coriolus versicolor Ps4a (Ohmomo et al., 1985b).

4. APPLICATION OF MICROBIAL-MDA FOR TREATMENT OF MWW 4.1. Application of Fungal Strain Many studies have been conducted on the decolorization process of MWW using fungal mycelia. For example, the mycelia of Coriolus versicolor Ps4a maintained a decolorization yield of about 75% for the continuous process under a dissolved oxygen (DO) concentration of 1 mg/L, dilution rate of 0.03 hr-1, and with the addition of 0.5% glucose and 0.05% peptone. This mycelia, immobilized in Ca-alginate gel, maintained the decolorization yield of about 65% for the continuous decolorization process under the dilution rate of 0.022 hr-1 for 16 days operation and removed about 53% of the COD value and 46% of the total carbon concentration, as shown in Figure 6 (Ohmomo et al., 1985b). Furthermore, the continuous decolorization of molasses wastewater by mycelia of Coriolus sp. No. 20 (Sirianuntapiboon and chairattanawan, 1998), the decolorization of MWW by the mycelia of Phanerochaetes chrysosporium immobilized into Ca-alginate gel (Fahy et al., 1997), the continuous decolorization of MWW by a mixture of immobilized-Coriolus versicolor IFO 30340 and Paecilomyces canadensis NC-1 strains (Terasawa et al., 2000), the decolorization of MWW by the mycelia of Coriolus hirsutus IFO 4917l in conjunction with the activated sludge process (Miyata et al., 2000) and the decolorization of MP from MWW by the immobilized mycelia of Flavodon flavus (Raghkumar et al., 2001) were also reported. In addition, under batch-type conditions, Mycelia sterilia D90 gave a maximum decolorization yield of about 80% and COD removal yield of about 70% in a three time replacement reaction for a continuous 24 day operation (Sirianuntapiboon et al., 1988b). Aspergillus fumigatus G-2-6 immobilized into a Ca-alginate gel maintained a decolorization yield of about 60% in the continuous replacement reaction for 18 days of operation and more than 70% in the continuous decolorization process under the dilution rate of 0.014 hr-1 with removal yields of about 51% (BOD5) and 56% (COD) (Ohmomo et al., 1987a). Aspergillus niger 180 maintained a decolorization yield of about 40% and a simultaneous COD removal

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yield of about 70% in the continuously fed batch system with MWW (Miranda et al., 1996). The decolorization of wastewater from alcohol fermentation using beat molasses was tested by using a process that combined the mycelia of Penicillium decumbens with the mycelia of Penicillium lignorum and Aspergillus niger, which resulted in a maximum decolorization yield of about 70% and a simultaneous COD removal yield of about 50% (Antonia et al., 2003). Also, Fujila et al (2000) also reported that polyurethane foam-immobilized white root fungi could be applied into the bioreactor for treatment of the molasses wastewater

4.2. Application of Bacteria Lactobacillus hilgardii W-NS immobilized into Ca-alginate gel gave about a 35% decolorization yield on the 7 times replacement decolorization process (for 30 days cultivation) and about 30% decolorization yield for continuous process under the dilution rate of 0.02 hr-1 and no aeration for 16 days, as shown in Figure 7 (Ohmomo et al., 1987b). Acetogenic bacteria BP 103 gave the decolorization yield of about 70% on the 6 times continuous replacement reaction for 30 days operation; however, it maintained only a 30% decolorization yield on the continuously fed batch reaction for 30 days operation (Sirianuntapiboon et al., 2004). Pseudomonas fluorescens adsorbed onto cellulose and coated by collagen gave about a 94% decolorization yield for the continuous replacement decolorization process (for 4 days operation) (Jagroop et al., 2001). A two-step column bioreactor using two bacteria strains of Pseudomonas putida U and Acetomonas sp. Ema was applied for the decolorization of molasses wastewater (Ghosh et al., 2002). The results showed that the COD and color intensity of MWW were reduced by 44.4% and 60.0%, respectively, in the first step by Pseudomonas putida U. Then the COD of the effluent of the first reactor was reduced by 44.4% in the secondary step with Acetomonas sp. Ema.

4.3. Application of Yeast Strain The continuous feeding system for decolorization of MWW by Citeromyces sp. WR-43-6 was tested and obtained a stable MP, COD and BOD5 removal efficiencies of about 50 - 60%, 99% and 89%, respectively, as shown in Figure 8. (Sirianuntapiboon et al., 2004b).

5. DISCUSSION MP is recognized as a material difficult to decompose (MDD) because it is hardly removed (decolorized or adsorbed) by normal biological treatment processes, such as the activated sludge system, aerated lagoons, anaerobic ponds, etc (Boopathy, 1992; Boopathy and Tilche, 1991; Sirianuntapiboon et al, 1988b). This review has outlined the potential application for microbiological removal of MP from MWW. Actually, the MDA has been detected among various groups of microorganisms, in spite of the classification of MP as an MDD. Although the removal or decolorization of MP by microbial processes is still only done on a laboratory-scale level, it is evident from this review that the microbial-MDA can be a useful process to treat, at least to decolorize, MWW. Among the microbial-MDA, the

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potential ability of lactic acid bacteria (LAB) should be useful for the development of the treatment process because LAB is a facultative bacteria and the treatment process requires noaeration. Immobilized LAB cells should be particularly advantageous because the treatment process is very simple and it is expected to be a low cost operation. And organic acids were generated was the raw material for the metanogenic bacteria group in the methane fermentation step of anaerobic treatment process. However, the MDA of LAB has only been detected on Lactobacillus hilgardii WN-S (Ohmomo et al., 1987c) and the decolorization yield was not very high when compared to some fugal strains, such as Coriolus versicolor Ps4a (Ohmomo et al., 1985b) and Aspergillus fumigatus G-2-6 (Ohmomo et al., 1987a). The screening of LAB having higher decolorization yields is expected. The fungal-MDA mentioned above and the yeast cells have a significantly higher decolorization yield. However, these abilities are unsuitable for treating MWW because these microorganisms are aerobic microorganisms and require oxygen for growth and MDA. Nevertheless, microbial-MDA has great potential for decolorizing MP. It is hoped that the MDA, such as those discussed in this review, will be exploited to their maximum potential in the near future. Also, the review article was consisted mostly our research group activity, we believed that the use of microbiological process to treat MWW was more suitable according to low cost, reusable of the end product and intermediate. It is therefore, recommended that further research regarding the MDA mechanism (both biological adsorption and degradation mechanisms) be conducted to further advance the understanding of biological MP removal. Also, the observation of some operation parameters of activated sludge systems (both aerobic and anaerobic process) such as dilution rate, organic loading, MP loading, sludge age and so on was necessary for the application of the potential-isolated strain in the conventional wastewater treatment processes.

Continuous decolorization was carried out in a bubbling column (diameter 26 mm x length 400 mm) at 37 °C. The column contained 100 ml of waste water and 60 g of immobilized cells mycelia (corresponding to 13.2 mg dry mycelial weight). Continuous decolorization was started by the feeding of wastewater after decolorization for 4 days (showed ↑and↓). The feeding rate was adjusted to 2.0 ml/hr (dilution rate of 0.02 hr-1). Symbols: , wastewater adjusted to pH 5.0 by NaOH; , wastewater adjusted to pH 7.3 by Ca(OH)2. Figure 7. Continuous decolorization of molasses waste water by immobilized cells of Lactobacillus hirgardii W-NS (Ohmomo et al., 1987c).

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Decolorization was carried out in an μ-carrier magnetic stirrer containing 10 ml of cells (4 x 109 cells/ml) and 1 liter of waste water added 2.0% of glucose, 0.1% of NaNO3 and 0.1% of KH2PO4 (pH 6.0) at 30℃ at an impeller speed of 150 rpm. From after 8 days culture, 100 ml of fresh medium was added into the system everyday. ↓ shows the time for feeding 10% of fresh medium. Symbols: , decolorization yield; , reducing sugar; , medium pH; ▲, dry cell weight. Figure 8. Decolorization in continuous feed system of molasses wastewater by Citeromyces sp.WR-436 (Sirianuntapiboon et al., 2004b).

6. CONCLUSION The microbial-MP removal process is the most suitable way to treat wastewater from factories that use molasses as a raw material. However, it is still only carried out on a laboratory scale. Many types of microorganisms, such as fungi (class Basidiomycetes and Ascomycetes and Dueteromycetes) and bacteria (Lactic acid bacteria and Acetogenic bacteria) were found to have MDA. The mechanisms of microbial-MDA might be the adsorption of MP on to the cell (cell membrane and cytoplasm) and/or the adsorbed or adsorbed MP was degraded by both intracellular and extracellular enzymes. Some of the microbes showed MP adsorption ability as the main activity, but others showed both MPadsorption and MP-adsorption activities. MDA was both induced and inhibited by the level of MP. Also, the aerobic, facultative, and anaerobic conditions of the microbial-MP removal processes were investigated. Rhizoctonia sp D-90 and Coriolus sp. No.20 were applied to the conventional wastewater treatment processes to treat MWW under aerobic conditions. Lactic acid bacteria was also introduced the anaerobic wastewater treatment process to decolorize MWW. Both Acetogenic bacteria BP 103 and Cyteromyces sp. WR-43-6 showed MDA with MWW under both aerobic and facultative conditions. Suspended growth (batch type and continuous type) and attached growth (Bio-film reactor) systems were tested for the microbial-MDA process for the treatment of molasses wastewater. However, all the experiments were still in the laboratory scale.

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Jagroop, D., Dalel, S. & Poonam, N. (2001). Decolourisation of molasses wastewater by cells of Pseudomonas fluorescens immobilised on porous cellulose carrier, Bioresource Technology, Vol. 78(1), 111-114 Jimenez, A. M, Borja, R. & Martin, A. (2003). Aerobic-anaerobic biodegradation of beet molasses alcoholic fermentation wastewater, Process Biochemistry, 38, 1275-1284. Kalavathin, D. F., Uma, L. & Subramanian, G.. (2001). Degradation and metabolization of the pigment–melanoidin in distillery effluent by the marine cyanobacterium Oscillatoria boryana BDU 92181. Enzyme and Microbial Technol., 29, 246-251. Kato, H. & Hayase, F. (2002). An approach to estimate the chemical structure of melanoidins, International Congress Series, 1245, 3-7(38). Kim, J. S. & Shoda, M. (1999). Decolorization of molasses and a dye by a newly isolated strain of the fungus Geotrichum candidum Dec 1, Biotechnology and Bioengineering, 62(1), 114-119. Kumar, P. & Chandra, R. (2006). Decolourisation and detoxification of synthetic molasses melanoidins by individual and mixed cultures of Bacillus spp. Bioresource Technology, 97(16), 2096-2102. Kumar, V., Wati, L., Banat, I.M., Yadav B.S., Singh D. & Marchant R. (1998). Decolorization and biodegradation of anaerobically digested sugarcane molasses spent wash effluent from biomethanation plants by white-rot fungi, Process Biochemistry, 33(1), 83-88. Kumar, V., Wati, L., Nigam, P., Banat, I. M., McMullan, G., Singh D. & Marchant, R. (1997). Microbial decolorization and bioremediation of anaerobically digested molasses spent wash effluent by anaerobic bacteria cultures. Microbios, 89, 81-90. Kwak, E. J., Lee, Y. S., Murata, M., Homma, S. (2005). Effect of pH control on the intermediates and melanoidins of nonenzymatic browning reaction LebensmittelWissenschaft und-Technologie, 38(1), 1-6. Lee, T. H., Aoki, H., Sugano, Y. & Shoda M. (2000). Effect of molasses on the production and activity of dye-decolorizing peroxidase from Geotricum candidum, J. Biosci. Bioeng., 89, 545-549. Miranda, M. P., Benito, G. G.., Cristobal, N. S. & Nieto C. H. (1996). Color elimination from molasses wastewater by Aspergillus niger, Biores. Technol., 57, 229-235. Miyata, N., Iwahori, K. & Fujita M. (1998). Manganese-independent and–dependent decolorization of melanoidin by extracellular hydrogen peroxide and peroxidase from Coriolus hirstus pellets. J. Ferment. Bioeng., 85, 550-553. Miyata, N., Mori, T. Iwahori, K. & Fujita, M. (2000). Microbial decolorization of melanoidin-containing wastewaters: Combined use of activated sludge and the fungus Coriolus hirsutus, Journal of Bioscience and Bioengineering, 89(2), 145-150. Mohana, S ; Desai, C. & Madamwar, D. (2007). Biodegradation and decolourization of anaerobically treated distillery spent wash by a novel bacterial consortium, Bioresource Technology, 98(2), 333-33. Monica, C., Garcia, M. T., Gonzalez, G., Pena, G. & Garcia, J. A. (2004). Study of colored components formed in sugar beet processing, Food Chemistry, 86(3), 421-433. Nakajima-Kambe, T., Shimomura, M., Nomura, N., Chanpornpong, T. & Takahara, T. (1999). Decolorization of molasses wastewater by Bacillus sp. under thermophilic and anaerobic conditions. J. Biosci. Bioeng., 87, 119-121.

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Ohmomo, S. Aoshima, I. Tozawa Y. & Ueda K. (1985b). Detection of lactic acid and amino acids from melanoidin decolorized by enzymes of Coriolus versicolor Ps4a. Agric. Biol. Chem., 49, 2767-2768. Ohmomo, S., Aoshima, I., Tozawa, Y., Sakurada, N. & Ueda, K. (1985a). Purification and some properties of melanoidin decolorizing enzymes, P-III and P-IV, from mycelia of Coriolus versicolor Ps4a. Agric. Biol. Chem., 49, 2047-2053. Ohmomo, S., Daengsubha, W., Yoshikawa, H., Yui, M., Kozaki, K., Nakajima, T. & Nakamura, I. (1988a). Screening of anaerobic bacteria with the ability to decolorize molasses melanoidin, Agric. Biol. Chem., 52, 2429-2435. Ohmomo, S., Itoh, N., Watanabe, Y., Aoshima, I., Tozawa Y. & Ueda K. (1985c). Continuous decolorization of molasses wastewater with mycelia of Coriolus versicolor Ps4a, Agric. Biol. Chem., 49, 2551-2555. Ohmomo, S., Kainuma, M., Kmimura, K., Sirianuntapiboon, S., Aoshima, I. & Atthasampunna, P. (1988b). Adsorption of melanoidin to the mycelia of Aspergillus oryzae Y-2-32, Agric. Biol. Chem., 52, 381-386. Ohmomo, S., Kaneko, Y., Sirianuntapiboon, S., Somchai, P., Atthasampunna P. & Nakamura, I. (1987a). Decolorization of molasses wastewater by a thermophilic strain, Aspergillus fumigatus G-2-6, Agric. Biol. Chem., 51, 3339-3346. Ohmomo, S., Yoshikawa, H., Nozaki, K., Nakajima, T., Daengsubha, W. & Nakamura, I. (1987b). Continuous decolorization of molasses wastewater using immobilized Lactobacillus hilgardii Cells, Agric. Biol. Chem., 52, 2437-2441. Raghukumar, C., Mohandass, C., Kamat, S. & Shailaja, M. S. (2004). Simultaneous detoxification and decolorization of molasses spent wash by the immobilized white-rot fungus Flavodon flavus isolated from a marine habitat, Enzyme and Microbial Technology, 35, 197-202. Raghukumar, C. & Rivonkar, G. (2001). Decolorization of molasses spent wash by the whiterod fungus Flavodon flavus, isolated from a marine habitat, Appl. Microbial Biotechnol., 55, 510-514. Sirianuntapiboon, S., Phothilangka, P. & Ohmomo, S. (2004a). Decolorization of molasses wastewater by a strain No. BP103 of acetogenic bacteria, Journal of Bioresource Technology, 92, 31-39. Sirianuntapiboon, S., Sihanonth, P., Somchai, P., Atthasampunna P. & Hayashida S. (1995). An absorption mechanism for the decolorization of melanoidin by Rhizoctonia sp. D-90. Biosci. Biotech. Biolchem., 59, 1185-1189. Sirianuntapiboon, S., Somchai, P., Ohmomo S. & Atthasampunna P. (1988a). Screening of filamentous fungi having the ability to decolorize molasses pigments. Agric. Biol.Chem., 51, 387-392. Sirianuntapiboon, S., Somchai, P., Sihanonth, P., Atthasampunna, P. & Ohmomo, S. (1988b). Microbial decolorization of molasses waste water by Mycelia sterilia D90. Agric. Biol. Chem., 52, 393-398. Sirianuntapiboon, S., Zohhsalam, P. & Ohmomo, S. (2004b). Decolorization of molasses wastewater by Citeromyces sp. WR-43-6, Process Biochem., 39, 917-924. Sirianuntapiboon, S. & Chairattanawan, K. (1998). Some properties of Coriolus sp. No. 20 for removal of color substances from molasses wastewater, Thammasart Int. J. Sc. Tech., 3, 74-78.

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Tamaki, H., Umetani, I., Tsuchida, H., Komoto, M., Arita I. & Hiratsuka, N. (1985). Decolorization of cane sugar molasses by action of Basidiomycetes (Part 1) - Screening of Basidiomycetes having high decolorization activity, Mem. Grad. School Sci. and Technol., Kobe Univ., A-3, 63-70. Terasawa, N., Murata, M. & Homm S. (2000). Decolorization of brown pigment in foods by immobilized mycelia of Coriolus versicolor IFO 30340 and Paecilomyces canadensis NC-1, J. Food Sci., 65, 870-875. Underkofler, L. A. & Hickely, J. (1954). Alcohol Fermentation of Molasses Industrial Fermentation, industrial Fermentation; Chemical Publishing Company, New York, 1-20. Watanabe, Y., Sugi, R., Tanaka, Y. & Hayashida, H. (1982). Enzymatic decolorization of melanoidin by Coriolus sp. No. 20, Agric. Biol. Chem., 46, 4623-1630. Yaylayan, V. A. & Kaminsky, E. (1998). Isolation and structural analysis of maillard polymers: caramel and melanoidin formation in glycine/glucose model system, Food Chemistry, 63(1), 25-31.

In: Fluid Waste Disposal Editor: Kay W. Canton, pp. 197-217

ISBN: 978-1-60741-915-0 © 2010 Nova Science Publishers, Inc.

Chapter 8

WASTEWATERS FROM OLIVE OIL INDUSTRY: CHARACTERIZATION AND TREATMENT

1

L. Nieto Martínez1, Gassan Hodaifa*2, Mª Eugenia Martínez1 and Sebastián Sánchez3

University of Granada, Department of Chemical Engineering, 18071 Granada, Spain. 2 Complutense University of Madrid, Department of Chemical Engineering, 28040 Madrid, Spain. 3 University of Jaén, Department of Chemical, Environmental and Materials Engineering, 23071 Jaén, Spain.

ABSTRACT Countries in the Mediterranean basin are among the main producers of olive oil. The elaboration of olive-oil is typically carried out by small companies in small facilities. The olive-oil plants produce high and variable amounts of residual waters of olives and oliveoil washing (OMW) that has a great impact in the environment. According to the procedure used different types of OMW with different chemical oxygen demand can be generated, the OMW from the three phase process (COD = 150 g O2 L-1) and the OMW from olives washing (COD = 0.8-4.5 g O2 L-1) and olive oil washing (COD = 1.1- 6 g O2 L-1) in the two-phase process. The uncontrolled disposal of OMW is a serious environmental problem, due to its high organic load, and because of its high content of microbial growth-inhibiting compounds, such as phenolic compounds. The improper disposal of OMW to the environment or to domestic wastewater treatment plants is prohibited due to its toxicity to microorganisms, and also because of its potential threat to surface and groundwater. These waters normally are stored in great rafts of accumulation for their evaporation during the summer. This solution among others until the moment dose not represent a definitive solution for this problem, especially as the administrations more and more demanding the preparation of this spill and the constructive quality of the rafts. Today, effective technologies have been proposed such as the chemical oxidation process using ferric chloride catalyst for the activation of H2O2 as a treatment of OMW produced from two-phase process. In the previous works the authors have described the *

Corresponding author. E-mail: [email protected] (G. Hodaifa), Tel.: +34 913 944 115; Fax: +34 913 944 114.

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L. Nieto Martínez, Gassan Hodaifa, Mª Eugenia Martínez et al. experimental results on laboratory-scale. These results have been taken to pilot-industrial scale, making the chemical oxidation in the optimum conditions of operations: [H2O2] = 5% (w/v), using a ferric chloride catalyst with a relation of [FeCl3]/[H2O2] = 0.25 (w/w), at OMW pH and environmental temperature. The final average value of COD obtained next to 370 mg L-1 (%CODremoval = 86.2%), and the water obtained can be destined to irrigation or disposed directly to the municipal wastewater system for their tertiary treatment. OMW from three-phase process does not allow direct chemical and biological purification for its content in phenolic compounds and generally used natural and forced evaporation process. Another way of using is the application of OMW nutrients to the growth of microorganisms such as microalgae.

1. INTRODUCTION Olive oil extraction produces vast amounts of liquid and solid wastes. The elimination of olive mill wastewater (OMW) is one of the main environmental problems related to the olive oil industry in Mediterranean countries, where Spain and Italy are the greatest producers. The OMW was genrated during a few months of the year (November-February). This liquid waste comes from the vegetable water of the fruit and the water used in the different steps of oil production and contains olive pulp, mucilage, pectin, oil, etc. suspended in a relatively stable emulsion. Olive oil is obtained by the traditional method of discontinuous pressing or by the continuous centrifugation of a mixture of milled olives and hot water. In both systems three phases are produced: (i) olive oil; (ii) solid by-product (olive pomace); and (iii) aqueous liquor, which represent 20, 30 and 50%, respectively, of the total weight of processed olives. The disposal of highly pollutant olive by products, especially the aqueous liquor, is an important environmental problem which needs to be solved. The aqueous liquor comes from the vegetation water and the soft tissues of the fruits. The mixture of this by-product with the water used in the different stages of oil elaboration constitutes olive mill wastewater (alpechin in Spanish). The quantity of OMW produced in the process ranges from 0.5 to 2 L kg-1 of olive, depending on the oil extraction system. In the main olive-oil-producing countries the implementation of systems based on the olive-pomace centrifugation has become more widespread. These include three- and twophase centrifugation systems. Effluents from two-phase systems are composed essentially by olive oil and olive-mill wastewater. The olive-oil and the wastewater must be separated. Continuous three phase extraction systems are still widely used in olive oil mills, especially in Italy, where in most cases they have not yet been replaced by more recent twophases systems, which involve a reduction of OMW volumes but an increased concentration in organic matter [1]. Three phases extraction systems involve the addition of large amounts of water (up to 50 L/100 kg olive paste), resulting in the worldwide production of more than 30 millions m3 per year of OMW [2]. This represents a great environmental problem, since this by-product is characterized by a high organic load; among the different organic substances found in OMW, including sugars, tannins, phenolic compounds, polyalcohols, pectins and lipids [3]. The toxicity, the antimicrobial activity and the consequent difficult biological degradation of OMW are mainly due to the phenolic fraction [4]. The partition coefficients (oil/water) of most olive phenols, ranging from 6 10-4 to 1.5 are in fact in favour of the water phase: the olive fruit is very rich in phenolic compounds, but only 2% of

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the total phenolic content of the olive fruit passes in the oil phase, while the remaining amount is lost in the OMW (approximately 53%) and in the pomace (approximately 45%) [5]. On the other hand, the phenolic compounds, which are very abundant in the OMW and are the major responsible of their polluting load, are characterized by a strong antioxidant activity [6]. The production of olive oil generates large volumes of wastes that vary in composition depending on which of the three production systems is used. The traditional method, which is based on the combined use of a crusher and hydraulic press, does not require the addition of water and yields a very high-quality olive oil. However, the system presents significant disadvantages to bulk production such as elevated labour requirements and a discontinuous process. This system has almost disappeared in many production areas. The three phase ‗continuous system‘, on the other hand, has many advantages, such as low labour costs and continuous production. However, it has disadvantages in the amount of wastewater produced. In some countries this system has almost disappeared due to the introduction of the two-phase decanter system. The two-phase decanter system is the newest system and is able to operate without water and thus dramatically reduces processing costs and the amount of wastewater produced. However, the semi-solid cake produced by this method has a high moisture content (55– 60%). One of the main disadvantages of the waste from 2- and 3-phase decanter systems is the presence of polyphenolic compounds in both the cake and the vegetation water. The presence of polyphenols limits the use of cake in animal food, since it can cause digestion problems in cattle [7]. Polyphenols also have been identified as being responsible for damage to soil when used for irrigation since they inhibit the growth of soil microflora [8]. The semi-solid waste from the two-phase decanter system cannot be purified by traditional methods so alternative solutions such as incineration and composting have been adopted for its treatment [9]. To solve the problems associated with OMW, different elimination methods have been proposed based on evaporation ponds, thermal concentration and physical-chemical and biological treatments, as well as its direct application to agricultural soils as an organic fertilizer. However, the most frequently used methods nowadays are the direct application to agricultural soils and storage in evaporation ponds, which produces a sludge. The disposal of OMW is becoming a critical problem in the Mediterranean countries. Traditionally the olive oil production sector was made up of a large number of small mills widespread throughout the production area. The volume of OMW produced in each mill was very small and its disposal was very widespread. These effluents rarely reached the water courses and their negative effects were only noticed in those places close to the mills. During the 1950s the industrialization of the olive oil production sector started, with the concentration of small producers in co-operatives and the creation of big factories with high milling capacities. Larger mills meant a greater local concentration and volume of OMW, which was discharged into the rivers without treatment. For this reason in 1981 the Spanish Government prohibited the discharge of OMW into the rivers and subsidized the construction of ponds for its storage during the milling period and the evaporation of its water during the warm Andalusian summer. Chemical oxidation based on Fenton‘s reagent (hydrogen peroxide in the presence of ferric salts) has been used to decompose organic and inorganic compounds in the laboratory, and also real effluents from several sources such as the textile and chemical industry, or

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refineries [10]. The process is based on the formation of various oxidizing agents which degrade pollution in wastewaters, but the nature of these species is under discussion [11,12]. Assays by Fenton at the end of the 19th century demonstrated that hydrogen peroxide and ferrous-salt solutions could oxidize tartaric and malic acids as well as other organic compounds. Later, Haber and Weiss [13] suggested that reaction (1). These radicals could react via the oxidation reaction) or via the attack on organic matter:

Fe2

Fe3

H2O2

Fe2

HO HO ; k 76Lmol-1 s-1

Fe3

HO

RH HO H2O

OH was formed through the of Fe2 to Fe3 (unproductive

(2)

HO

ROH H3O

(1)

oxidizedproducts

(3)

At pH < 3.0, the reaction is autocatalytic, since the Fe3+ decomposes H2O2 to O2 and H2O through a chain reaction:

Fe3

Fe OOH2

H2O2

Fe OOH2 Fe2

H2O2

H

HO2 Fe2

HO2 Fe3 HO H2O2

(5)

HO HO

(6)

HO2

(7)

O2 H

(8)

H2O HO2

(9)

Fe3

HO2 Fe2

(4)

Fe3

Fe2

The process is potentially useful to remove pollutants, as it is very effective in generating

HO . However, an excessive amount of Fe2 ions could trap or consume them (reaction 2), as happens with halogens, H2O2, or the radical perhydroxyl:

HO HO2

O2 H2O

(10)

Today, it is thought that other species of Fe(IV) or Fe(V) (such as FeO2 and ferrule complexes), are the actual active agents of the process. In the presence of peroxide, the Fe2+ concentration is low compared to the Fe3+concentration, since the reaction (4) is slower than reaction (6). Both radicals, HO and HO2 , react with organic matter, but HO2 is less reactive. The speed constant for the reaction of ferrous ion with H2O2 is high, and Fe(II) is oxidized to Fe(III) after a few seconds or minutes in excess H2O2. Therefore waste destruction through Fenton‘s reagent is thought to be a process catalysed only by Fe(III)– H2O2 and Fenton‘s reaction with an excess of H2O2 is essentially a process of Fe3+/H2O2. For

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this reason these kinds of reactions also occur with transition metal ions such as Fe(III) or Cu(II), and they are known as Fenton-type reactions:

Men

H2O2

Me(n 1)

HO HO (Me = Fe3+, Cu2+)

(11)

As a general rule, the degree and total mineralization speed are independent of the initial oxidation state of the Fe. However, the initial efficiency and speed of demineralization are higher when starting from Fe(II). On the contrary, Fe(III) salts produce an stationary Fe(II) concentration. In this case a pH lower than 2.8 must be used. Fenton‘s process has been effective for the degradation of aliphatic compounds and aromatic chlorates, PCBs, nitro aromatics, azoic colorants, chlorobenzene, phenols, chlorate phenols, chlorates, octachloro-p-dioxin, and formaldehyde. Only a few compounds cannot be attacked by this reagent, such as acetone, acetic acid, oxalic acid, paraffins, and organochlorinated compounds. This reagent is a good oxidizer for herbicides and other soil pollutants such as hexadecane or Dieldrin. It is used to decompose dry-cleaning solvents and to decolour wastewater containing different kinds of colorants and other industrial wastes, reducing its COD. Fenton‘s reaction has been successfully applied to reduce the COD of municipal and underground water and also for the treatment of lixiviates. It is highly useful for the pre-treatment of non-biodegradable compounds. The advantages of this method are the following: Fe2+ is an abundant non-toxic substance, and hydrogen peroxide is easy to handle and environmentally friendly. Unlike other oxidative techniques, chlorinated compounds do not result and there are no masstransfer limitations because the system is homogeneous. The reactor design for the technological application is very simple. However, it requires continuous and stoichiometric addition of Fe2+ and H2O2 as well as a high Fe concentration. It is important to take into account that excessive Fe2+ amounts may cause the proper conditions for HO to be trapped, according to the above-mentioned equations. At pH > 5.0 particulate Fe3+ is generated, although this produces sludges demanding later management, and, at the end of the process, water is usually alkalinized by means of simultaneous addition of flocculants to remove waste iron. The molar stoichiometric H2O2/substrate reaction should theoretically oscillate between 2 and 10 if a reagent is used for the destruction of soluble compounds. In practice, however, this relation may reach values of up to 1000, since the destructible compound of many environmental samples is usually accompanied by other compounds that can be attacked by HO . The peroxide/Fe/compound relation may be maintained by intermittently adding oxidizer or be fixed at the beginning of the reaction. In the laboratory, the metal aggregate is traditionally made in the form of pure ferrous salts, but high prices of these salts hamper their use at the industrial level. Instead of salts, Fe2(NH4)2SO4 containing 20% active iron is used. Other iron compounds have been used, including some solids such as goethite to remove trichloroethylene [14]. In such cases total mineralization is not reached, but some intermediaries resistant to the treatment are created (carboxylic acids), that react slowly with HO , and there is a predomination of the unproductive reaction where ferrous ion is converted to ferric ion. More toxic products than the initial ones, such as quinones, can sometimes form, and these must be thoroughly controlled.

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Rivas et al., [15] have recently studied the oxidation of p-hydroxybenzoic acid (PHB) with Fenton‘s reagent, a pollutant usually found in effluents generated by the food industry. An optimal molar relation Fe/H2O2/pHB of around 5.10-3/2.65/1.10-2 was established. The addition of tert-butyl alcohol, an HO trap, had very little influence on the process, and therefore it is interpreted that other radicals were present, as mentioned above. The formation of phenol, catechol, hydroquinone and trihydroxybenzene indicates that a degradation mechanism acted via decarboxylation. They also studied the action on atrazine, its derivatives, and other pesticides. Microalgae contribute to sustainability in environmental conservation by the photosynthetic fixation of CO2 from atmosphere and gaseous industrial effluents, as well as through the consumption of different C, N, and P compounds in urban and industrial wastewaters. These actions of microalgae, either naturally or induced by humans, are possible due to the different nutritional modes presented by most microalgae. The opposite situations are autotrophic and heterotrophic nutrition. Diluted OMW from the three-phase system can be used as nutrient medium for the growth of Scenedesmus obliquus. This study report the characterization of different wastewaters from modern olive oil industry. Concretely, the use of Fenton‘s reaction and microalgae to treatment the wastewater from the two and three phase process, respectively.

2. EXPERIMENTAL 2.1. Wastewater In the Andalusian provinces of Jaén and Córdoba (Spain) wastewater samples were collected from several oil mills operating with olive-cleaning and vertical-centrifugation equipment of different trademarks.

2.2. OMW Treatment by Fenton’s Reaction After the analysis and characterization of the samples, one with high chemical-oxygen demand was collected, and a mixture of olives and olive-oil wastewaters in a proportion of 1:1 (v/v) was prepared in the laboratory. The pH, electric conductivity, and COD were determined. Chemical oxidation based on the Fenton‘s reagent was used for the treatment of the effluents generated by olive-oil mills operating with a continuous two-phase system. Optimal operating conditions at room temperature included the hydrogen peroxide concentration, as well as the catalyzer concentration and type, and the coagulant concentration was identified previously [16,17]. The best catalyst ferric chloride (efficiency and low cost) from among different compounds was chosen. The COD value of water at the end of the chemical oxidation at different concentrations (5, 7.5, 10, and 30% w/v) was determined for each catalyst, maintaining a relationship [catalyst]/[H2O2] = 0.05 (w/v).

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The [catalyst]/[H2O2] (w/w) was varied between 0.1 and 1.5 with the aim of determining the optimum value for the maximum degradation of organic matter and phenol compounds. The oxidation process was completed with pH-neutralization and separation phases (solidliquid) to produce the irrigation water. The experiments were made in a stirred tank reactor. During the experiments, the pH, temperature, and electric conductivity values were determined in relation to time. 500 mL of wastewater with a COD = 7.2 g O2/L, electric conductivity = 1.52 mS/cm and a pH = 4.6 was mixed with the appropriate amount of ferric catalyst and H2O2. The catalyst and H2O2 dissolution were added gradually during the course of oxidation. The mixture was stirred for 2.5 h, which is enough time to complete the oxidation. The reaction is exothermic only when a catalyst is used. The solid phase and the liquid phase were separated by decanting. The final COD in the liquid phase was determined and from this value the reduction percentage in this parameter was calculated.

2.3. OMW Treatment by Microalgae The freshwater microalga used was Scenedesmus obliquus CCAP 276/3A, supplied by Culture Centre for Algae and Protozoa, Oban (United Kingdom). The experiments were performed in stirred batch tank reactors on a laboratory scale. The photobioreactors, 5 total, were situated in a culturing facility described elsewhere [18,19]. Each reactor had 0.75 L capacity (70 mm in diameter and 200 mm in height) with thermostatically controlled water circulation, magnetic stirring, and aeration. The culture medium wastewater was prepared with ultrapure water (Millipore, mod. Milli-Qplus) for concentrations of 2.5%, 5%, 10% and 20% OMW (v/v). The pH was adjusted to an initial value of 7.0 and maintained over the course of the culture. The working temperature was 25ºC. All the cultures were mechanically stirred at 350 rpm and supplied air sterilized by filtration (0.2 m pore size), at a specific rate of 1 v/v/min. The illumination was continuous, at an intensity of 298 E m-2 s-1 (QSL 2100, Biospherical Instruments, Inc.). The mean value of the initial biomass concentration was 0.0124 g L-1 and the standard deviation SD = 0.0080 and the initial cells number was 0.315 109 cell mL-1 and SD = 0.197 109.

3. RESULTS 3.1. Characterization of OMW 3.1.1. OMW from two phase process The physical-chemical composition of wastewater from the olive-oil industry changed during the collecting campaign for several reasons, especially for the type of olive variety, geographic location, form of collection and if olive soil or not. In the Table 1 are reflected the values for the parameters analyzed in some of the wastewaters of examined washer machines of olives. It is observed that almost all the waters

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practically fulfil the values demanded in the normative one (The waters should not overcome the values for the following parameters: pH = 6-9, suspended solids = 600 mg L-1, BOD5 = 1.000 mg O2 L-1, COD = 1.000 mg O2 L-1, Spain legislation). Only one of them has an inferior pH to 6 units. The suspended solids are always inferior to 600 mg L-1. Only three samples overcome allowed COD and BOD5. It is deduced that, in general, most of the present total solids is from mineral character to the being the superior percentage of ashes to the percentage of organic matter. Only in a case flotation is detected, that is to say a layer floating whose composition denotes that it is probably and reasonable oil of the sweat of broken or damaged olives. A treatment of chemical oxidation, with energetic oxidizers, with a later alkaline correction of pH followed by a sedimentation or filtration would be enough, to adapt the water to the demanded requirements. After a short period of natural sedimentation, it would lose a part of the small fraction of suspended solids and it could be clever for its use in watering. In the Table 2 are reflected the values of the parameters analyzed for wastewater of vertical centrifuges of olive oil of the same almazaras for those that we are took waters of olive laundry. Contrary to that exposed previously for the waters of olive washers, in this case all the samples overcome the reference values, to exception of the suspended solids since is not detected sedimentation neither separation of phases. The muddiness of the samples should be caused by the fine emulsion of oil in water. In all the cases the COD is bigger than the allowed values. As expected, in many cases the quantity of total solids is smaller than in the wastewater of olive washers and in this case the percentage of the organic matter is bigger than the percentage of the mineral matter. As it was already exposed in the introduction the biggest quantity in organic matter of these waters it is due to the own composition of the oil and they should be rich in phenolic compounds, natural and recalcitrant antioxidants to their microbial degradation and therefore their concentration will be reflected difficultly in the figures of BOD5 and if on the contrary in those of COD. It seems therefore that this water is the main wastewater to try. A priori it has been thought of subjecting it in the first place and for their sour character to processes treatment of chemical oxidation, with energetic oxidizers, H2O2 with iron salts (Fenton treatment). The first experiments carried out lead to an important decrease of the colour and the COD, as well as to an increase of the pH whose value is inside the allowed limits (pH < 6). Table 1. Parameters of wastewater of olive washers. Sample 1 (Co) 2 (Co) 3 (Co) 4 (Co) 5 (J) 6 (Co)

pH 6.34 5.65 6.22 6.66 6.02 6.03

Total solids (%) 0.27 0.49 0.23 0.18 0.28 0.87

Ashes (%) 0.17 0.27 0.07 0.08 0.21 0.53

Organic matter (%) 0.10 0.22 0.15 0.10 0.07 0.34

Suspended solids (%) n.s.f n.s.f 0.005 0.006 0.006 n.s.f

Sediment Solids (%) n.d. n.d. 0.225 0.174 0.274 n.d.(Flot.)

BOD5 mg L-

COD mg L-

500 1820 348 148 121 1145

810 4858 1640 223 810 4494

1

1

n.s.f = Separation of phases is not observed after one hour; n.d = Sedimentation is not detected; Flot. = Flotation is observed with a superficial layer. CO = Cordoba (Spain); J = Jaén (Spain).

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Table 2. Parameters of wastewater of vertical centrifuges. Sample 1 (Co) 2 (Co) 3 (Co) 4 (Co) 5 (J) 6 (Co)

pH 5.69 5.40 5.67 5.73 5.11 5.16

Total solids (%) 0.18 0.15 0.24 0.33 1.47 0.59

Ashes (%) 0.04 0.05 0.04 0.07 0.05 0.10

Organic matter (%) 0.14 0.10 0.20 0.26 1.42 0.49

Suspended solids (%) n.s.f. n.s.f n.s.f n.s.f n.s.f n.s.f

Sediment Solids (%) n.d n.d n.d n.d n.d n.d

BOD5 mg L-

COD mg L-

790 520 465 690 915 790

2875 5935 3806 4231 12078 10931

1

1

n.s.f = Separation of phases is not observed after one hour; n.d = Sedimentation is not detected; Flot. = Flotation is observed with a superficial layer CO = Cordoba (Spain); J = Jaén (Spain).

Table 3. Parameters of wastewater from three phase process. Parameters pH Moisture (%) Total solids (%) Organic substances (%) Fats (%) Ash (%) BOD5 (g O2 L-1) COD (g O2 L-1) Conductivity (mS cm-1) % C (% dry matter) % N (% dry matter) % H (% dry matter) % O (% dry matter) % S (% dry matter) % P*** % K*** Polyphenols (% dry matter)*** Carbohydrates (% dry matter)***

Raw OMW* 5.4 93.4 6.6 5.8 1.54 0.9 42.0 151.4 7.9 50.9 1.4 7.1 40.5 0.1 0.19 5.24 2.21 12.2

Filtered OMW* 5.6 95.4 4.6 3.9 0.8 n.a** n.a** 7.1 42.0 1.1 5.3 51.5 -

*OMW, Wastewater from the three phase process of olive oil extraction (3 phases: Solid + Liquid + Liquid) **n.a., not available ***Reported by Paredes et al. (1999)

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3.1.2. OMW from three phase process The three-phase process, still used in many olive-oil extraction mills, generates a residual effluent (OMW) that has a high organic load. OMW, a dark brown wastewater, contains vegetable water from the olive fruit itself, from the washing of the fruit, from the washing of the olive oil, and from other activities in the mill. This wastewater is markedly acidic. Although the BOD5 and COD values are 21 and 13 fold greater than those found in the residual waters produced by the two-phase process, respectively. The filtration of the raw wastewater decreased the C and N content by 17.5 and 21.4%, respectively. OMW has a high content in total solids, reaching 66 g L-1 compared to 5 g L-1 for OMW from two phase (Table 3) and 1.2 g L-1 for untreated, highly loaded urban sewage. Its fatty content accounts for 1.54%. Table 3 also lists the contents P, K, polyphenols, and carbohydrates provided by Paredes et al. [20]. The ratios H:C, N:C, O:C, and P:C of the OMW were 0.13, 0.026, 1.22 and 0.005, respectively, while the ratios of microorganism biomass according to the elemental formula of Harrison [21] C H1.64 N0.16 O0.52 S0.0046 P0.0054 were 0.14, 0.19, 0.69, and 0.014, indicating a certain deficiency in N and P in the wastewater. The ratio N:P of the OMW and biomass 5.8:1, 13.6:1 reflected a larger N deficit.

3.2. OMW Treatment by Chemical Oxidation (Fenton’s Reaction) In all the experiments the values of pH, electrical conductivity and temperature was determined with the course of the chemical oxidation reaction (Figure 1). A temperature increase was detected in all the experiments, due to the strong exothermal reaction. This increase over the ambient temperature was determined for the different peroxide concentrations used. Electrical conductivity also increased with the course of the oxidation reaction, this increase is logical considering that the amount of catalyst added (Fe source) increased during the experiment. Moreover, it is important to point out the decrease in the pH to values around 3.0, which is an optimal value for Fenton‘s reaction (Figure 1).

0

pH, Conductivity (mS/cm) and T ( C)

45 40 35 30 8 6 4 2 0 0

20

40

60

Time (min)

Figure 1 The values of pH, electric conductivity, and temperature variation with the course of chemical oxidation reaction (□ pH, ● electric conductivity and ∆ temperature) maintaining a rate [catalyst]/[H2O2] = 0.05 w/v, catalyst used Fe(NH4)2(SO4)6H2O, temperature 25 ºC, [H2O2]initial = 5% w/v.

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Table 4. Final values of COD, purification efficiency and sediment solids determined for the water treated by chemical oxidation reaction (initial conditions: COD = 7.2 g O2/L, electric conductivity = 1.5 mS/cm, T = 25 ºC). CODFinal (g O2/L)

Ferric chloride

5.0 7.5 10.0 30.0

1.27 1.46 1.44 2.38

Purification Efficiency (% COD) 82.3 79.7 80.0 76.0

4000

12

3500 COD (mg O2/L)

Sediment solids v/v 0.80 -

10

3000 8

2500 2000

6

1500

4

1000 2

500 0 0.0

0.2

0.4

0.6

0.8

1.0

total phenols (mg/L)

[H2O2] % (w/v)

[Fe] or

Catalyst applied

0 1.2

FeCl3/H2O2 (w/w)

Figure 2. Relationship between the values for COD (■), [Fe] (▲) and total phenols (○) of treated water and the [FeCl3]/[H2O2] relationship. Operating conditions: initial values of COD = 4104 mg O2/L, total phenols = 290 mg/L, and [Fe] = 5 mg/L.

Table 4 shows the degradation results expressed as the final values of COD and the depuration efficiency (reduction percentage in the final COD) for ferric chloride salt. Also, Table 4 shows the sediments solid determined after the neutralization of water treated and separation test in Imhoff cone during 1 h. The best concentrations of hydrogen peroxide to work was varied in the range from 5 to 10% w/v, where the depuration efficiency has varied between 76 and 82%. The highest value of sediment solids (0.8 v/v) has been obtained. For the determination of the best catalyst/H2O2 relation, a series of experiments was performed using ferric chloride. The relation was varied between 0.1 and 1.5. Figure 2 shows a downward trend of all the parameters for the values of the Fe/H2O2 relation between 0.25 and 0.5 (w/w). For this reason, the following experiments were performed with Fe/H2O2 values of 0.25, in order to reduce catalyst consumption. From Figure 1, it can be deduced that the pH fell from 4.0-5.0 of the treated water to values of around 3.0, at which the reaction occurs optimally. For this reason, a neutralization process of the oxidized water was performed in order to adjust the pH to neutrality, as the use regulations demand. At the same time, the iron ion found was removed as hydroxide, which is difficult to precipitate. For the precipitation, several flocculants were tested and an oil-based

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anionic one from the Nalco Company was chosen. For the determination of the concentration needed, several experiments were performed to determine the influence of the flocculant concentration with relation to the time necessary for settlement in an Imhoff cone, the COD, and the remaining amount of Fe in the treated water. The resulting values were indicated no significant influence on the COD, while the Fe concentration decreased to a concentration of 1 mg/L and, from this value, the concentration remained stable. As mentioned above, the experiment was conducted at a pilot-plant scale (3-5 m3 h-1) in S.A.T. Olea-Andaluza olive-oil mill factory in Baeza, Jaén (Spain). In this olive-oil mill factory, function and verification tests of the results found at the laboratory scale were performed. Only the environmental conditions changed, because, although the plant was roofed in, it was exposed to the elements (environmental temperature between 1-7ºC during the mornings). The plant worked intermittently during the 2004/05 harvest and subsequently was automated in order to work continuously for the 2005/06 season. The process in the plant consisted of: 1. Natural sedimentation in independent holding pools for water from olives and olive-oil washing; 2. Chemical oxidation tank, 3. Neutralization tank and coagulant addition; 4. Separation of solid and liquid unite by decanting; 5. Filtering unite. As explained above, the plant worked with mixtures of approximately 1:1 (v/v) of wastewater from olives and olive-oil washing. Table 5 shows the values of the parameters analysed in all the streams. The working conditions except the temperature (which was the ambient temperature) were deduced in the laboratory: a relation of [FeCl3]/[H2O2] of 0.25 (w/w), neutralizing agent NaOH, and coagulant concentration 1 mg/L. Some 4-5 m3 per charge were used intermittently, the oxidation time being 2.5 h. The final filtration was performed first with sand and afterwards with a biomass filter (olives stone). Also, Table 5 shows the same parameters for oxide water at the outlet of the reactor. A COD value decrease of about 61% and the total phenolic compound content of about 68% can be deduced. At the same time, there was an expected increase in electrical conductivity caused by the addition of a catalyst; part of that increase was eliminated in the stage of decantation and filtration. At the exist of pilot plant practically all the phenolic compounds were removed, and the Fe decreased by about 78% and COD value by about 76%. Independently of what happened in the neutralization (Fe(OH)3 precipitation), which in principle should not carry organic matter, these decreases were due to the filtration process used, in which the biomass fill of one of the filters (olives stone) acted as an adsorbent of heavy metals. The water obtained was taken to a waterproof storage pool for use in irrigation. Today, the plant is computerized for continuous functioning and its treatment capacity is about 3 m3 h-1 that is, enough to treat the wastewater produced by an olive-oil mill factory with a capacity of approximately 600.000 kg olives day-1, which corresponds to a medium/large olive-oil mill factory.

3.3. OMW from Three Phase Process Treatment by Microalgae For all the experiments, growth curves showed no lag phases, the first phase being exponential growth followed by a phase of linear increase in biomass with time. In some experiments, the stationary phase was observed, and even the onset of cell death.

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Table 5. Characterization of industrial wastewaters from olives and olive-oil washing, wastewater in the oxidation reactor, waters at the exit of oxidation reactor and at the outlet of pilot plant. Wastewater in the oxidation reactor* 6.10 2.25 2763 282.6 7.74 555 7.2

pH Conductivity, mS/cm COD, mg O2/L Total phenols, mg/L [Fe], mg/L [Cl], mg/L [H2O2]residual, mg/L Total pesticides, µg L1

Water at the exit of oxidation reactor 7.0 4.02 1077 90.1 5.87 716 615 n.d.**

Water treated at the exit of pilot plant 7.4 3.91 661 0.005 1.67 834 n.d.**

*This wastewater formed mixing the wastewaters from olives and olive-oil washing with the ratio of 1 v/v.

The specific growth rates, μ = d(lnx)/dt, during the exponential-growth phases, μm, was calculated according to Eq. (12)

ln

x x0

(12)

a μm t

were plotted against the initial OMW concentration So, expressed in % (v/v), Figure 3. 0.06

%OMW (v/v) WC RL WF UW

0.05

0.03

m

h

0.04

0.02 0.01 0.00 0

5

10

15

20

25

% [S0], v/v

Figure 3. Variation of the maximum specific growth rates with the initial concentration of OMW (WC: OMW without color, RL: Rodríguez-López medium, WF: OMW without fatty matter, UW: urban wastewater from secondary treatment as the medium culture). Common conditions: aeration 1 v/v/min, agitation speed 0 350 rpm and illumination intensity = 298 E m-2 s-1.

The variation of μm with So, appears to indicate an inhibitory effect in the wastewater. This was to be expected, as the OMW may contain fats, organic acids, phenolic compounds, and the remains of pesticides, which are known to harm microalgal growth [22].

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From the different inhibition models by substrate and toxicity assayed, the one that best reproduced the experimental variation observed was the Teissier [23] one of inhibition by substrate, Eq. (13), solid line in Figure 3. Table 6. Reduction in the contaminant load of the OMW. [S0] (%OMW v/v) 2.5 5 10 20 WF WC

x-x0 (g L-1) 0.0156 0.0419 0.0876 0.0577 0.0506 0.113

BOD5, initial (mg O2 L-1) 1100 2200 4400 8800 2100 2100

BOD5, end (mg O2 L-1) 190 1650 1900 3100 849 950

BOD5,removal (mg O2 L-1) 83 550 2500 5700 1251 1150

Initial concentration of the biomass at t = 0 h was 0.0124 g L-1.

μm μm, max e

S0 Ki

-

e

S0 Ks

%BOD5,removal 82.7 25.0 56.8 64.8 59.6 54.8

BOD5, removal/x-x0 (g g-1) 5.3 13.1 28.5 98.8 24.7 10.2

(13)

Though the values μm, max = 0.032 h-1 ([S0] = 10 v/v), Ki = 87%, and Ks = 2.83% are consistent with what observed. At low initial OMW concentration the S. obliquus has a high affinity for the limiting quantity of the substrate, resulting in a low Ks value. Roughly speaking it is the division between the lower concentration range where μm is strongly (linearly) dependent on S0, and the higher range, where μm becomes independent of S0. The high value of Ki (87% v/v) indicated that the inhibition effect can be observed only in a high concentration range (cultures with OMW > 10%). All the cultures received the same aeration level (1 v/v/min), agitation velocity and the illumination intensity kept the same, 298 E m-2 s-1, but the attenuation of the light, by the coloration of the medium, was greater the higher the %OMW, and thus the variation expected, in μm , being light the limiting factor. This fact was confirmed in the control

experiment WC (OMW without color) where the μm value was increased to 0.04 h-1 for the same culture concentration (5% v/v). But the main factor limiting growth was the fat matter, where the value of μm registering for WF experiment (OMW without fatty matter) was 0.05 h1

similar to that determined for the mineral medium RL. This can be explained may be determine the greater distortion in the composition of the biomass, increasing the fatty matter percentage with the augment of %OMW in the culture, of an oily nature, can be adsorbed onto the cell surface, hampering the access of nutrients. However, in the culture formed with urban wastewater (UW) it has been determined 0.052 h-1 for μm slightly major than for mineral medium (RL). On the other hand, at the end of the cultures, it was determined the contaminant load for the wastewater after separation from the biomass (Table 6). A greater net reduction of BOD5 was achieved in cultures with 20% OMW (5.7 g O2 L-1). But the higher %BOD5, removed was determined in the culture with 2.5% OMW. The decline in BOD5 shows the use of organic compounds by the microalga. The maximum biomass generated values and BOD5, removed are

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detected in the culture with 10 and 20% OMW, respectively. This circumstance at 20% OMW may determine the greater distortion in the composition of the biomass, as components of the undiluted wastewater, of an oily nature, can be adsorbed onto the cell surface, hampering the access of nutrients (especially O2). The control experiments WF and WC are registering an increments in the net biomass generation and in the %BOD5, removed (Table 6). This confirms the effect of an oily nature medium in access nutrients to the cells and the importance of light in the net biomass generation.

4. CONCLUSIONS OMW from Two Phase Process Treatment by Fenton Reaction’s The chemical oxidation (Fenton reaction) studied in this work are able to treat olive mill wastewater precedent from two phases continuous system. The best removal efficient was achieved for COD (76%). The phenolic compounds were destroyed. Treated water resulting from this process is used to irrigate. This process offers a solution for reducing the environmental effect of wastewaters generated by two-phase centrifugation system of oliveoil industry. The catalyst used (FeCl3) containing ferric iron ions which lead to savings in the consumption of oxidizer (to avoid rust ions Fe II to III). The sediments obtained in the decanter are dredging mud creamy rich in iron. Water obtained are a fully transparent without odors, phenolic compounds or pesticides.

OMW from Three Phase Process Treatment by Microalgae Obviously, this wastewater without pre-treatment is not an appropriate medium for the cultivation of the microalga S. obliquus. The strong inhibition of growth during the exponential phase and nitrogen deficiency necessitate a pre-treatment prior.

REFERENCES [1] [2] [3] [4]

Roig, A., Cayuela, M. L. & Sánchez-Monedero, M. A. (2006). An overview on olive mill wastes and their valorisation methods. Waste Management, 26, 960-969. Borja, R., Alba, J. & Banks, C. J. (1997). Impact of the main phenolic compounds of olive mill wastewaters (OMW) on the kinetics of acetoclastic methanogenesis. Process Biochem., 32, 121-133. D‘Annibale, A., Crestini, C., Vinciguerra, V. & Giovannozzi Sermanni, G. (1998). The biodegradation of recalcitrant effluents from an olive mill by a whit-rot fungus. J. Biotechnol., 61, 209-218. Bisignano, G., Tomaino, A., Lo Cascio, R., Crisafi, G., Uccella, N. & Saija, A. (1999). On the in-vitro antimicrobial activity of oleuropein and hydroxytyrosol. J. Pharm. Pharmacol., 51, 971-974.

214 [5] [6] [7] [8] [9] [10] [11]

[12] [13] [14] [15] [16]

[17]

[18] [19] [20]

L. Nieto Martínez, Gassan Hodaifa, Mª Eugenia Martínez et al. Rodis, P. S., Karathanos, V. T. & Mantzavinos, A. (2002). Partitioning of olive oil antioxidants between oil and water phases. J. Agr. Food Chem., 50, 596-601. Obied, H. K., Allen, M. S., Bedgood, D. R., Prenzler, P. D. & Robards, K. (2005). Investigation of Australian olive mill waste for recovery of biophenols. J. Agr. Food Chem., 53, 9911-9920. Molina, E. & Nefzaoui, A. (1996). Recycling of olive oil by-products: possibilities of utilization on animal nutrition. Int. Biodeterior. Biodegrad., 38, 227-235. Cappasso, R., Evidente, A. & Schivo, L. (1995). Antibacterial polyphenols from olive mill wastewaters. J. Appl. Bacteriol., 79, 393-398. Tomati, U., Galli, E., Fiorelli, F. & Pasetti, L. (1996). Fertilisers from composting of olive mill wastewaters. Int. Biodeterior. Biodegrad., 38, 155-162. Bigda, R. J. (1996). Fenton‘s chemistry: an effective advanced oxidation process. J. Adv. Sci. Eng., 6(3), 34-37. Bossmann, S. H., Oliveros, E., Göb, S., Siegwart, S., Dahlen, E. P., Payawan, L., Straub, M. Jr., Wörner, M. & Braun A. M. (1998). New evidence against hydroxyl radicals as reactive intermediates in the thermal and photochemically enhanced Fenton reactions. J. Phys. Chem. A., 102, 5542-5546. Pignatello, J. J., Liu, D. & Huston, P. (1999). Evidence for additional oxidant in the photoassisted Fenton reaction. Environ. Sci. Technol., 33, 1832-1836. Haber, F. & Weiss, J. (1934). The catalytic decomposition of hydrogen peroxide by iron salts. Proc. Roy. Soc., 147, 332-351. Teel, A. L., Warberg, C. R., Atkinson, D. A. & Watts, R. J. (2001). Comparison of mineral and soluble iron Fenton's catalysts for the treatment of trichloroethylene. Wat. Res., 35(4), 977-984. Rivas, F. R., Beltrán, F. J., Frades, J. & Buxeda, P. (2001). Oxidation of phydroxybenzoic acid by Fenton's reagent. Wat. Res., 35(2), 387-396. Martínez-Nieto, L., Rodríguez, S., Giménez, J. A., Lozano, J. L., Cobo, A., Ortega, J. & Hodaifa, G. (2003). Efluentes de la industria del aceite de oliva: contribución al estudio de la composición y tratamiento de las aguas de lavado de aceituna y de lavado de aceite. In: Estudio de la composición y tratamiento como subproducto de las aguas de lavado de aceituna y aceite, 13-44, Ed. Infaoliva, Granada, Spain. Martínez-Nieto, L., Rodríguez, S., Giménez, J. A., Lozano, J. L., Cobo, A. & Hodaifa, G. (2004). Procesos oxidativos en el tratamiento de las aguas de lavado de aceituna y de lavado de aceite. In: Aguas de lavado de aceituna y aceite: procesos de tratamiento, 73-102, Ed. Infaoliva, Córdoba, Spain. Hodaifa, G. (2004). Aprovechamiento de las aguas residuales de la industria oleícola en la producción de biomasa de microalgas. PhD Thesis, University of Jaén, Faculty of Experimental Science, Jaén, Spain. Sánchez, S., Martínez, M. E. & Espínola, F. (2000). Biomass production and variability of the microalga Isochrysis galbana in relation to culture medium. Biochem Eng J, 6, 13-18 . Paredes, C., Cegarra, J., Roig, A., Sánchez-Monedero, M. A. & Bernal, M. P. (1999). Characterisation of olive mill wastewater (alpechin) and its sludge for agricultural purposes. Bioresource Technol., 67, 111-115.

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[21] Harrison, J. S. (1967). Aspects of commercial yeast production. Process Biochem., 2, 41-45. [22] Kobayashi, H. & Rittmann, B. E. (1982). Microbail removal of hazardous organic compounds. Environ. Sci. Technol., 16, 170-181. [23] Teissier, G. (1936). Les lois quantitatives de la croissance. Ann. Physiol-Chim. Biol., 12, 527-573.

In: Fluid Waste Disposal Editor: Kay W. Canton, pp. 215-239

ISBN: 978-1-60741-915-0 © 2010 Nova Science Publishers, Inc.

Chapter 9

USABILITY OF BORON DOPED DIAMOND ELECTRODES IN THE FIELD OF WASTE WATER TREATMENT AND TAP WATER DISINFECTION

(a)

Hannes Menapace(a*), Stefan Weiß(b), Markus Fellerer(a), Martin Treschnitzer(a) and Josef Adam(a) Institute for Sustainable Waste Management and Technology - University of Leoben, Franz-Josef-Strasse 18, 8700 Leoben, Austria. (b) Umweltbundesamt GmbH, Spittelauer Laende 5, 1090 Vienna, Austria.

ABSTRACT Over the past few years one main focus on the research efforts at the Institute for Sustainable Waste Management and Technology (IAE) has been on possible applications for reactors with boron doped diamond electrodes (BDD) in the field of (waste) water treatment. This article deals with the technical construction of the electrodes used (continuous reactor with a different number of plate electrodes), which were produced by a spin-off of the institute. The electrodes consist of conductible industrial diamond particles (< 250 µm), which are mechanically implanted on a fluoride plastic substrate. These electrodes showed a high mechanical and chemical stability in different test runs. At the institute, treatment methods for micro pollutants (e.g. pharmaceuticals and complexing agents) were developed with electrochemical oxidation by BDD. In this case test runs were made on laboratory scale and technical scale treatment units and elimination rates up to 99 % were achieved. In this project the analytic is partly provided by the ―Umweltbundesamt GmbH‖ (UBA), one of the project partners. This agency has been a project partner in different studies about pharmaceuticals in the ecosystem. These techniques could also be used for the waste water treatment of alpine cabins. Pilot projects have been set up. On the basis of these results a follow-up project was launched last October, in which an alternative treatment process for oil-in-water emulsions and mixtures was developed by the usage of electrochemical oxidation with BDD. A third possible application is the disinfection of drinking water from contaminated ground and *

Corresponding author: E-mail: [email protected], Phone: +43 3842 402 5105, Fax: +43 3842 402 5102.

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Hannes Menapace, Stefan Weiß, Markus Fellerer et al. spring water. In this process oxidation agents like ozone or OH radicals produced in situ by the BDD reactor from the treated water are used to eliminate bacterial contaminants (for example e. coli) in the water.

1. DIAMOND ELECTRODES Due to their mechanical and chemical stability boron doped diamond electrodes are well suited for the treatment of fluid waste and media. During the so called electrochemical advanced oxidation process oxidizing agents are directly produced out of the organic matrix of the treated fluid. The chemical structure of the organic matrix and pollutants are degraded by these agents and the chemical oxygen demand (COD) will be decreased. For instance the double bonds in the chemical structure of the pollutant are split up and functional groups are cracked. Thus biodegradability is increased. Based on experimental research at various universities, several spin-offs have been established since the year 2000. These companies conduct research, distribute diamond electrodes, diamond coatings for applications in the field of water and waste water treatment. In Europe the distribution of such electrodes and reactors is dominated by three companies.

1.1. Producers and Different Electrode Types The following chapter gives an overview on three main producers of boron doped diamond electrodes (BDD-electrodes) regarding treatment of fluid waste.

1.1.1. Adamant Adamant Technologies SA was founded 2005 in Switzerland. It is a spin-off company of CSEM, Centre Suisse d‘Electronique et de Microtechnique S.A.. The production facility is located in the Science and Technology Park Neode, in La Chaux-de-Fonds (NE). The fields of activities lay mainly in Diamond Coating Technology. Regarding water treatment applications the so-called Adamant®-Electrodes and complete systems (DiaCell®Technology) are available. Additionally, the company is active in water process monitoring. The diamond coatings are produced by chemical vapor deposition technique (CVD). [1] 1.1.2. Condias The CONDIAS GmbH is a spin-off of the Fraunhofer Institute for Thin Films and Surface Technology. The company was founded in 2001 and has its headquarters in Itzehoe, near Hamburg, Germany. The main products are diamond electrodes with the trade name DIACHEM®. These electrodes, also produced by the chemical vapor deposition technique (CVD) on different base materials like Nb, Ta, Ti, Graphite, Si or conductive ceramics, are primarily used for waste water treatment and electrochemical synthesis. CONDIAS produces diamond coated areas up to 100 x 50 cm² with diamond layer thickness up to 15 µm. [2] 1.1.3. Pro aqua The pro aqua GmbH is a spin-off company of the University of Leoben. It was founded 2002 and has its headquarter in Niklasdorf, Austria. Primarily the company produces and sells

Usability of Boron Doped Diamond Electrodes…

219

boron doped diamond electrodes with a layer of titanium oxide. In contrast to the other two producers of BDD electrodes mentioned, diamond particles up to a size of 250 µm are mechanically implanted on the metal substrate. In 2006 a new electrode type was developed and patented. The old model of the titanium substrate was replaced by a film of fluorinated plastic. The company distributes smaller flow rate reactors for waste water treatment and disinfection of supply and tap water, but also standard BDD electrodes with a maximum area of 16 cm x 16 cm. Greater electrode areas are supplied by a special welding methode [3]

1.2. Construction and Design of BDD Electrodes In the following chapter the construction of BDD electrodes is explained on the basis of the pro aqua patent concerning the bipolar diamond electrodes with a substrate of fluorinated plastic. This type of electrodes was used for the different degradation tests which were conducted on the Institute for Sustainable Waste Management and Technology (IAE) at the University of Leoben, Austria. Furthermore, the advantages and disadvantages of characteristic settings are discussed.

1.2.1. Comparison of Different Electrode Types During the first few years of the electrode development titanium was used as the coating material. Several techniques for application of the doped diamond particles on a titanium layer were investigated. Technical problems occurred in all investigations. For example the so called passivation effect, the formation of an oxide layer between the boundary layer of the diamond particle and the titanium, could not be prevented. This effect leads to a continuous decrease of the active surface of the electrode. As a result, the production rate of the oxidizing agents and the treatment efficiency decreases. A second problem is caused by different material expansion coefficients of diamond and titanium. Cracking fissures on the electrode surface result in the chipping of the diamond particles. Due to these technical problems, the concept for the pro aqua electrode was completely reviewed in 2006 and fluorinated plastic was chosen as coating material in 2007 [4]. Neither a passivation effect nor the described cracking fissures arose. This electrode type shows a wide range of application due to its mechanical and chemical stability. Therefore it has been successfully used in several research projects, carried out by the Institute. The invention of the pro aqua GmbH [4] relates to a method of producing a diamond electrode comprising synthetically produced and electrically conductive (doped) diamond particles (2) which are embedded in a support layer (1) made of electrically nonconductive material. The doped diamond particles (2) are introduced as a single layer between two films that form the support layer (1). The films are then permanently connected to each other, and the diamond particles are exposed on both sides of the support layer (1). A scheme of the objective electrode is shown in Figure 1.

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Figure 1. Diamond electrode comprising a support layer of electrically non-conductive material [4]

Table 1. Currently available electrode types [1,2,3,4]. Property Types Dopant Available current density Coating methods Layer materials Feeding electrodes Interelectrode distance Active Area Pretreatment Polarity reversal frequency

Available Settings Monopolar and bipolar Boron In the research on the IAE densities up to 100 mA/cm² were applied. The value varies depending on application Chemical vapor deposition (CVD), mechanical implementation Si, Nb, TI, Graphite, conductive ceramics Cu, Al, Pt, mixing oxides 2 – 10 mm up to 5.000 cm² Filtration recommended 1 min to 90 min

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1.2.2. Comparison of Different Electrode Types A short overview of the currently available electrode types is shown in the following Table 1. Furthermore, some characteristic settings are explained. Interelectrode distance vs. conductivity Depending on the particular application different gaps between the electrode plates in the treatment reactor are needed. Especially fluids with a high content of suspended solids require interelectrode distances. Otherwise fouling would be the result. The greater the gap between the electrodes the higher is the potential drop per electrode plate. Hence, for the treatment of fluids with a low conductivity (< 500 mS/cm) a smaller gap is recommended. Polarity change Through an automatic polarity changer lime precipitation on the electrodes will be avoided. Hereby it‘s guaranteed that the active electrode area is constant during the treatment process.

1.3. Advanced Oxidation Process with Diamond Electrodes In the field of waste water treatment and fluid waste disposal, the anodic oxidation process is a rather new technology, which is not widely used. This treatment method falls into the category of electrochemical oxidation processes and is an ideal additional treatment step for conventional disposal systems, especially if no biodegradable substances should be treated or in presence of toxic chemicals which would inhibit biodegradation processes. [5] The chemicals commonly used for this purpose are oxygen, hydrogen peroxide, ozone, permanganate or persulfate. The higher the oxidation potential of the reagent used, the more efficient the chemical oxidation process is. The most powerful oxidant in water is the hydroxyl radical with a redox potential of 2.8 V relating to normal hydrogen electrode (VNHE) [6]. Organic contaminants are degraded into inorganic substances such as H2O, CO2, and the waste water is additionally disinfected by the agents produced. In general the higher the oxidation potential, the higher the efficiency of the treatment process. Table 2 shows the oxidation potential of some chemical substances. Table 2. Oxidation potential of some oxidation agents [6]. el.-chem. Potential (V)

Potential rel. Chlorine (%)

Oxidation agent

Symbol

Fluorine

F2

3.06

2.25

Hydroxyl radical

*OH

2.80

2.05

Oxygen atom

*O

2.42

1.78

Ozone

O3

2.08

1.52

Hydrogen peroxide

H2O2

1.78

1.30

Hypochlorous acid

HClO

1.49

1.10

Chlorine

Cl2

1.36

1.00

Chlorine dioxide

ClO2

1.27

0.93

Oxygen

O2

1.23

0.90

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Figure 2. Working range of different electrode materials relating to the potential of the standard hydrogen electrode [6].

The anodic oxidation can be ascribed to the Advanced Oxidation Process (AOP). This term comprises all oxidative methods, which have hydroxyl radicals as the main oxidation agent. Hydroxyl radicals act as the main oxidation agent. As can be seen in the table, the OH.radical with an oxidation potential of 2.8 V is a fairly strong oxidation agent in water. During the anodic oxidation process the boron doped diamond electrodes form the basis of the process. These synthetic diamond electrodes differ from other electrodes due to their high mechanical and chemical stability and the manner in which water electrolysis is carried out. Whereas during the electrolysis of water with conventional electrodes the water molecule is split into oxygen and hydrogen, during treatment with a diamond electrode, highly reactive hydroxyl radicals are formed. This is due to the high potential of a diamond electrode. Previous tests conducted with graphite or carbon bearing electrodes have shown that the use of these materials is not advantageous as they do not exhibit the necessary potential to produce hydroxyl radicals [5]. Moreover, signs of wear appear all too readily due to the fact that, in addition to the formation of oxygen, CO2 formation also occurs leading to the breakdown of the electrode material. Even electrodes comprised of PbO2, SnO2 or Pt do not show enough efficiency in the production of OH radicals and are not suiTable due to their low mechanical and chemical stability [5]. Figure 2 shows the range of activity of the different electrode materials vs. the hydrogen potential (left side), oxygen (right side). The potential required to create hydroxyl radicals lies at 2.8. In contrast, boron doped diamond electrodes work with an energy efficiency of more than 90%. This means that in the anodic oxidation process the OH radicals are formed in an electro-chemical manner directly from the waste water treated and the impurities are mineralized or at least transformed into biologically degradable materials, without producing any further residues or waste. The hydroxyl radicals formed react with the impurities in the waste water by splitting hydrogen. [6] H2O  OH + e- + H+

(1)

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Figure 3. Combined chlorine content depending on active electrode area and applied current density, UB untreated water sample, R3K 9 plates, R2K 6 plates and R1K 3 plates.

The most suitable alternative method is chemical oxidation, aiming for the total mineralization or the production of harmless or biodegradable compounds by use of oxidants. By using the electrochemical production of hydroxyl radicals, no additional chemical substances are necessary. The process can be performed at affordable costs, determined by the power required for driving the electrochemical process and without the common AOP drawbacks. Electrochemical water disinfection by producing disinfecting agents (mainly active chlorine produced from the naturally dissolved chloride ions) during electrolysis of water is another common water treatment process. The amount of electrochemical produced oxidizing agents (for example chlorine content shown in Figure 3) manly depends on the two parameters active area and current density. Figure 3 shows the content on combined chlorine for different reactor sizes. Therefore a different number of BDD electrode plates (active area per plate) were installed in each reactor. To achive a better efficiency of the process, a static mixer was implemented in downstream of the reactor R1K in the test run R1K SM.

2. FIELDS OF APPLICATION 2.1. Treatment of Pharmaceuticals and Complexing Agents in Waste Water 2.1.1. Introduction Pharmaceuticals are discharged into the sewer system with human or animal excrements and finally end up in the municipal sewage plant. Because of their complex chemical structure some drugs (e.g. Carbamazepine) or chelating agents (for example

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EDTA and NTA) cannot be eliminated using conventional waste water treatment procedures, so they pass into the aquatic system [7, 8]. As an example the release of pharmaceuticals into surface waters may lead to increased dissemination of antibiotic resistance [9], endocrine substances like hormones are suspected to promote feminizing effects on organisms in ecosystems [10]. Complexing agents like EDTA may cause a remobilization of sedimented heavy metals in surface waters. While there are already statutory thresholds for EDTA and NTA implemented in Austria (QZV Chemie OG 2006) [11] according to the EU directive 2000/60/EC [12], a regulation for pharmaceuticals is expected in the near future. To be able to meet these requirements two innovative treatment procedures have been designed. One is the anodic oxidation with boron doped diamond electrodes and the other one the ozonation. For the second method a new sort of ozone generator is used and the ozone is injected into the water flow in a venturi injector. The development of both procedures is in a testing phase. During research at the Institute for Sustainable Waste Management and Technology at the University of Leoben a new process design was developed in two steps. First a small lab unit was constructed (flow rates ranged from 380 L/h). Treatment sources included synthetic waste water, cleaned waste water from the local municipal waste water treatment plant and a wide range of sectoral waste water. The technical results from this first phase were used for the design and construction of the technical scale unit.

2.1.2. Theoretical process background The basic idea for the treatment of pharmaceuticals and industrial chemicals was to combine the anodic oxidation and the ozonation, as both process steps are able to provide the needed oxidants (O3 and Hydroxyl radicals). These two processes are quite different, so both are described briefly. When using boron doped diamonds electrodes, oxidants (O3 and hydroxyl radicals) are extracted directly from the waste water's organic matrix, which is done by applying direct current to the electrodes. These oxidants are able to eliminate the organic compounds thus no additional chemicals are needed. For ozonation, the oxidant (O3) was generated by electrical production at the beginning and subsequently is being produced by a sort of dielectric barrier discharge. 2.1.3. Description of equipment During the research project two technical plants were constructed and are being operated: a small plant in laboratory scale located at the institutes laboratory and a medium sized technical scale unit located at the local municipal waste water treatment plant. 2.1.3.1. Bench-scale unit During the first phase of the project, the lab-scale unit (Figure 4) was used to determine the process parameters relevant for further process design and to gain insight on the essential parts and sizes for constructing the technical scale unit. In the second step, the small plant was utilized for several test series treating different sources of sectoral waste water.

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Figure 4. Sketch of the of the laboratory scale unit - a combination of anodic oxidation and ozonisation.

The first laboratory scale unit (Figure 4) was used in the first project step to determine the process parameters relevant for further process design and to gain insight on essential parts and sizes for constructing the tech scale unit. In the second step, the small plant is utilized for several test series treating different sources of sectoral waste water from hospital and industry. The plant follows a modular design concept and consists of two independent segments, one represents a flow reactor for the anodic oxidation, the other a reaction well with the attached ozone generator. The parts operate either separately or in combination. Hydraulic conveyance of the waste water to a particular reaction circuit is provided by diaphragm pumps (Sera R203-2,4E 3 L/h) and flexible-tube pumps (Gardener, Watson-Marlow 323e max. 86 L/h).

2.1.3.1.1. Anodic Oxidation The treatment unit for the anodic oxidation in the laboratory scale consists of a flow reactor, which is equipped with eight parallel plate electrodes (total area 352 cm²). The treated medium seeps through this reactor, process parameters are determined by downstream sensors. As experiences gained in the preliminary tests pointed out, treatment improved when using a downstream catalyst based on metal oxide, so this type of catalyst was used for an after-reaction for the follow up test series. Voltage supply for the electrodes is provided by an EA-HV 9000-600-2000 power supply (I = 0 ~ 3 A), while the current density is kept at a constant level.

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2.1.3.1.2. Ozonisation The ozonisation process consists of two steps: the production of O3 and the treatment reactor for contacting the oxidant with the waste water. This has the advantage of there being two ways to optimize the treatment plant. In laboratory scale different sorts of production and insertion were investigated. At the beginning diamond electrodes located on titanium oxide plates were used, but as these caused technical problems (electrode lifetime, operating stability) the process was finally substituted by an advanced corona-discharge generator for the ozone. Reduction tests were carried out in counter current flow at the beginning, in subsequent project steps a venturi injector was used to mix waste water and ozone. Similar to the anodic oxidation, interconnection of sensors is possible. 2.1.4. Tech scale unit The technical scale unit (TSU) consists of four parallel waste water flows, three are equipped with electrochemical reactors, manufactured by "pro aqua Diamantelektroden GmbH", of different electrode areas (3, 6 and 9 electrodes per reactor). This plant design (Figure 5) permits a reference value on the fourth water flow and the possibility to take three samples at the same time for comparison. Furthermore, the unit is equipped with an automatic polarity changer to avoid lime precipitation on the electrodes. During the next stage of research, an automatic sample collector and a fluid level indicator for the waste water source were connected to the plant further increase automation. 2.1.5. Test series From January to September 2007 miscellaneous test series were carried out using the laboratory scale unit (LSU) on the Institute for Sustainable Waste Management and Technology (IAE) to examine the single procedure steps and combinations of the reactors used in waste water treatment process. To gain knowledge about the interaction of the different chemicals in the matrix and to optimize the treatment the fluid flow rate and the current density were varied. Based on these experiments, test series on the technical scale unit and series with sectoral waste water were started in the second project phase (Table. 2)

Figure 5. Tech scale unit; left side: controlling station, right side: reactor unit.

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Table 2. Overview of the test series. Projec t phase

I

II

Experiment parameter

Synthetic waste water with EDTA

Aggregate Laboratory scale Technica unit l scale unit Anodic Ozoni Oxidatio n sation x x

Degradability experiments with pharmaceutics endowment Real waste water without additional endowment Variation of current densities and flow rates Different contact methods Treatment combinations

x

x

x

x

x

x

x

x x

Experiments with industrial waste water Variation of current densities and flow rates Serial connections of the reactors

x x

x x

Venturi injector for the ozone contact Ozonization as reference method

x x

x x

2.1.6. Sample preparation For determination of so called principal parameters fixing agents have to be added immediately after sampling. Thus further degradation of ingredients is prevented. To prevent breakdown of pharmaceuticals 100 mg NaN3/L sample is added, to fixate chelating agents. Formaldehyde (w = 37 %, 10 mL/L) provided the capabilities needed. The samples then were sent to the UBA in cooling boxes. 2.1.7. Analytics The process parameters temperature, pH-value and redox potential were directly monitored by sensors and the anodic oxidation current and voltage were recorded. The analysis of pharmaceuticals and chelating agents (Table 3) was carried out by the 'Umweltbundesamt' (UBA), where similar projects had been carried out before [10, 13]. During the first preliminary tests of the two technologies applied on the LSU an analysis of the EDTA elimination was made. A complexometric titration according to DIN method DIN 38406-3 [14] for determination of calcium and magnesium ions in water by EDTA was used. Titrating with a calcium solution of defined concentration gives the possibility to calculate the amount of EDTA in the treated solution. For this EDTA was added to the untreated water sample. After treatment 50 mL samples were taken and sodium hydroxide (NaOH, 2 mol/L) and an indicator salt were added. The sample was then titrated with a calcium chloride solution (CaCl 2, 50 mg/L) until a colour

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change from blue to purple occurred. The concentration of EDTA in the titrated sample was then calculated according to the DIN standard DIN 38406-3.

Table 3. Analysis program. Substance groups Sum parameter

Pharmaceutical

Complexing agent

Experiment parameter

IA E

DOC

x

Redox potential, pHvalue, conductivity COD, mg/L Carbamazepine Caffeine Roxithromycine Erythromycine-H2O Josamycine Diazepame Trimethoprime Sulfamethoxazole EDTA, µg/L NTA, µg/L DTPA, µg/L 1,3-PDTA, µg/L

x

LOQ: Limits of Quantification, LOD Limits of Detection.

UB A

LOQ ng/L 10 mg/L ---

x x x x x x x x x x x x x

15 2.0 20 20 20 20 2.0 20 20 1 1 2.5 1

LO D ng/L ------1,0 10 10 10 10 1.0 10 10 1 1 5 1

2.1.7.1. Pharmaceuticals For analysis of pharmaceutical compounds 500 mL of the samples were acidified, spiked with an isotopically marked surrogate standard mixture and subsequently enriched by means of solid phase extraction. Analytes were eluted using dichlormethane, ethylacetate and methanol. The resulting extract was concentrated under a gentle stream of nitrogen and solvents were changed to acetonitrile and water. The final extract was spiked with an internal standard to follow instrument stability and compensate for matrix effects. Samples were analyzed by means of liquid chromatography-electrospray ionization-tandem mass spectrometry. Quantification was performed by external standard method. 2.1.7.2. Complexing agents For analysis of complexing agents isotopically marked surrogates and an internal standard were added to the samples. Samples were concentrated to dryness on a sand bath at 120 °C, and the residue was resolved in 1M hydrochloric acid. After evaporation of the acid the residues were esterified with a mixture of n-butanol and acetylchloride. After the reaction was stopped by addition of sodium hydroxide solution, the resulting esters were extracted with n-hexane and dried over sodium sulphate and finally analyzed with gas chromatography-

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mass spectrometry in Single Ion Recording (SIR) mode. Quantification was performed with internal standard method by means of isotope dilution.

2.1.8. Analysis During the experiments (Figure 6) we observed a strong dependency of the treatment success on the applied current density, the reactor surface and the time of contact could be observed. The results also showed a different degradability of the individual substances as Carbamazepine showed a better degradability than Diazepam. The degradability of Complexing agents showed a deterioration of the degrading performance at a low concentration range (µg/L-Area). 2.1.9. Summary With the experiments completed thus far, the applicability of the treatments used for a continuing waste water treatment was proven [15]. Furthermore, the experiments on the TSU were performed under the most realistic conditions possible to collect data material for the optimization (e.g. reactor dimension to increase the contact time). Based on a comparison of the success of a particular treatment of the municipal and industrial waste water, a statement concerning the applicability for central and decentral waste water treatment was made. In addition to the treatment success the costs of investment and the costs of treatment should be considered. Based on 0.07 €/kWh and after a first estimation for the flow rate with 200 L/h the costs for the treated waste water depending on current densities from 30.2 to 42.3 mA/cm² will range between 0.16 and 0.60 €/m³.

Figure 6. Actual rates of elimination for the Anti-epileptic drug Carbamazepine, TSU – different current densities and flow rates.

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2.2. Treatment of Oil-Water-Emulsions and Mixtures In all companies where waste is disposed of with state-of-the-art technology, a fat separator will be built in. The discharge of waste water into the sewer system should meet the legal parameters such as COD, temperature, pH value and the rate of lipophilic substances. However, the high content of lipophilic materials in waste water that has not be precleaned leads to fat depositions in the pipe system of the sewer system. Reduced pipe cross section, blockage and smell nuisances are the results. So an alternative secondary treatment step of the waste water from fat separators is of utmost priority. By means of an anodic oxidation process by BDD electrodes, oil water mixtures from a chemically - physical (CP) treatment system have been treated at the IAE in a research study since October 2008. the COD will be analyzed in the experiments. Furthermore, the parameters of conductivity, pH value and temperature will be continuously logged. Flow reactors which are equipped with bipolar pro aqua diamond electrodes are used as a main item for this electrochemical oxidation process. Feeding electrodes made of mixing oxide are used for the power supply of the reactor. To gain knowledge about the energy input degradation tests with different current density are carried out. The following degradation test carried out with an oil water mixture sample from a chemical physical treatment system. The sample was taken before the ultrafiltration step of the facility. In Table 4 the technical data for the used reactor is shown. Table 4. Technical data of the used flow reactor. proaqua-reactor-1: Diamond electrodes Active area per electrode Total area Gap between electrodes max. current density Feeding electrodes Housing material Seals Flow rate

4 plates 32,5 cm2 130 cm2 3 mm 50 mA/cm2 Ru/Ir coated titanium sheet Polypropylene Viton max. 50 L/h

Table 5. Technical data – power supply. Power supply EA-HV 9000-600-2000: Main AC voltage

230 V/AC

Max. output DC voltage

0 - 600 V/DC

Max. output DC current

0 – 3,3 A

Type

linear and adjustable

Reversion of polarity

0,05 Sec. - 24 h.. (adjustable)

In this case 2.000 mL of emulsion were treated in a batch-system with a flow reactor, applied with BDD diamond electrodes under a flow rate of 21.8 L/h. A current density of 83.3 mA/cm² was reached by an active electrode surface per plate of 4 x 30 cm² and a constant

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current of 2.5 A. The duration of the process was limited to 5 h 30 min. Figure 7 shows the resulting degredation effect by a current density of 83.3 mA/cm².

Figure 7. Degradation of COD for oil-water emulsion after ultra-filtration – 83.3 mA/cm², 21.8 L/h.

2.3. Treatment of Pesticides in Drinking Water Exemplified by Atrazine Contamination Atrazine, a prioric substance according to the European Water Framework Directive [12], is a herbicide which was used mainly for the weed control on sweet corn cultivation, vini- and pomicultire and along stretches of rail road. Along with the main metabolites desethylatrazine, desisopropylatrazine and hydroxyatrazine, atrazine is in the triazine group . Due to its effectiveness and in lieu of alternatives the inexpensive chemical agent was applied in great quantities and on a large scale. In 1994 the usage of atrazine was banned in Germany, one year later the herbicide was also banned in Austria according to the Plant Protection Act. On the European level the ban was adopted about ten years later in 2004. In 1991 an active agent amount of approximately 400,000 kg were placed into circulation, in the following years the consumption decreased rapidly down to 5,000 kg in 1995 [16]. These data were obtained by reason of notification requirement according the Austrian Plant Protection Act. The main areas of atrazine and desethylatrazine contamination are found in the agriculture regions of Southeastern and Eastern Austria.

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Figure 8. Concentration of atrazine and its metabolites in spring water treated with BDD electrodes.

Although atrazine displays only a marginal acute toxizity, studies showed critical effects in the field of body weight gain, inhibition of ovulation and effects on the heart function. Furthermore, the substance is suspected to have carcinogenic effects. Hence, for consumers an ADI value (Acceptable Daily Intake) of 0.005 mg/kg bw (body weight)/d (day) was defined. Moreover, endocrine effects were detected for atrazine. [17, 18, 19] Depending on the local factors such as rainfall, soil moisture and the adsorptive properties of the soil, atrazine finds its way into the groundwater. Due to its higher mobilization rate the risk of desethylatrazine getting into groundwater is higher in comparison to atrazine. [20] Within the scope of a smaller project, degradation tests of atrazine and disethylatrazine contaminated spring water were conducted at the Institute. In this case the unpurified water samples were taken from a spring in an area with intensive agricultural usage. For the treatment a flow reactor according the specifications of Table 4 was used. Figure 8 shows the degredation effect for atrazine and its metabolites at an applied current density of 45.5 mA/cm². After treatment (batch as well as continuous operation mode) of the herbicide contaminated spring water, the concentration of desethylatrazin was below the threshold according to the Austrian Drinking water directive.

2.4. Disinfection of Tap and Drinking Water In the alpine region of middle Europe thousands of alpine lodges and mountain inns were built in exposed positions. In view of the sensible ecosystem and in fact that many lodges were built above the catchment area of drinking water abstraction, untreated waste water

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disposal is forbidden. A biological step alone, such as a decrease of degradation performance at sinking temperature, is frequently not sufficient. In a project supported by the local Styrian government, the biological treatment step at a lodge (1,524 m above sea level) was upgraded with an additional disinfection system based on BDD electrodes. The scientific support of the optimization of these added treatment steps was supported by the IAE. In a second work in this sector, the bacteriological pollution of spring water was treated by anodic oxidation with BDD electrodes. A reactor cell from the pro aqua GmbH was used to support drinking water quality according to the specifications of the Drinking Water Ordinance. For both problems samples were taken and treated by anodic oxidation. The disinfection effect for waste water effluents with the objective reactor was verified by an expert opinion. In different studies successfully tests were made with BDD electrodes to inactivate legionella in water samples. [22, 23] Using BDD electrodes with direct current supply under galvanostatic conditions lead to an electrolytic generation of oxygen based disinfectants like peroxodisulfate, peroxodicarbonates, hydrogen peroxide and OH radicals on the active electrode area. These agents could be used to eliminate bacteria and other organic components. For the test runs two different arrangements were used. In addition to treatments in a batch operation, investigations with a continuously treatment were also tested. The used reactor is composed of bipolar four plate electrodes with an active area of 32.5 cm² per electrode and two feeding electrodes made of mixing oxide. During the tests a 2 mm gap between the BDD electrodes was used. A batch operation mode mainly is reasonable for the treatment of highly contaminated waste water when the contact time of the media in one stroke through the reactor is insufficient. For the validation of the treatment effect under laboratory scale conditions the following parameters were tested Escherichia coli Enterococci Coliformic germs Colony forming units (CFU) at 22 °C and 37 °C The treatment of the waste water from the alpine lodge shows the dependence between disinfection efficiency on the one hand and the parameters flow rate and current density on the other hand (Table 6). Sample 1: Untreated influent (8 °C water temperature; 2.5 mS/cm conductivity) Sample 2: effluent after the disinfection step at a pumping capacity of approx. 30%. Sample 3: effluent after the disinfection step at a pumping capacity of approx. 22%. Sample 4: effluent after the disinfection step at a pumping capacity of approx. 36%. In the test runs with the contaminated spring water, values for the treated water below the quantification limits could be achieved. According to the results a continuous operation mode of the reactor (current density of 36.92 mA/cm² and a flow rate of 43.5 L/h) seems sufficient for the treatment.

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unit

CFU at 22 °C CFU at 37 °C Enterococci

CFU/mL CFU/mL CFU/100 mL

Sample 1 93

Sample 2 30

Sample 3 1.7 x 104

Sample 4 1.9 x 104

4.8 x 104 0

675

9.8 x 103

9.9 x 103

0

0

509

Table 7. Microbiotic Parameters – DVGW-Analysis (unfiltrated waste water treatment plant effluent, temperature: 20 °C, sample volume: 2500 mL, flow rate: 50 L/h, Applied Voltage: 60 V; current: 0.5 A, treatment time: 100 min [24]. Parameter

Unit

Escherichia coli

CFU/100 mL CFU/100 mL CFU/100 mL CFU/100 mL

California germs Enterococci CFU (36°C)

Waste water effluent before treatment 2.807

Waste water effluent after treatment <1

15.600

<1

2.120

<1

76.970

<1

Table 7 shows microbiotic examinations of waste water effluent from the sewage water treatment plant in Karlsruhe, Germany before and after treatment by a flow reactor with the pro aqua diamond electrodes. The analysis was carried out by the Heinrich-SontheimerLaboratory for water technology of the DVGW (German Technical and Scientific Association for Gas and Water) in Karlsruhe, Germany [24].

3. CONCLUSION AND OUTLOOK An advantage of the anodic oxidation process with BDD electrodes is based on the fact that now usage of additional chemicals is needed for the treatment process. Due costs are generated by investments in the reactors and the power supply. Estimation of treatment cost is relatively simple. In the sector of drinking water purification, the anodic oxidation has been explored much better and is frequently used in practice. Thanks to the possible adaptation of electrode materials, from simple techniques to concrete problem statements, a treatment of different contaminated fluid wastes is appropriate. The usage of BDD electrodes is an effective technology as a secondary treatment step for disinfection. In this respect different research studies have shown the effect of inactivation. The fluid should exhibit a minimum conductivity for the electrochemical treatment.

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4. REFERENCES [1] [2] [3] [4] [5] [6] [7] [8] [9] [10]

[11] [12] [13]

[14] [15] [16] [17]

Adamant Technologies SA (2009). Specification sheets. http://www.adamantec.com/ CONDIAS GmbH (2009). Specification sheets. http://www.condias.de/. 2009 pro aqua GmbH (2009). Specification sheets. http://www.proaqua.cc. 2009 Schelch, M., Staber, W., Wesner, W., et al. (2007). Method for the production of a diamond electrode and diamond electrode, Patent: WO 2007/116004 A2/PCT/ EP2007/053337, 18. October. Kraft, A., Stadelmann, M., et.al. (2003). Anodic oxidation with doped diamond electrodes: a new advanced oxidation process. Journal of Hazardous Materials, 247261. Tröster, I., et.al. (1998). Electrochemical advanced oxidation process for water treatment using DiaChem electrodes. Diamond and Related Materials, 640-645. Ternes, T. (1998). Occurrence of drugs in sewage treatment plants and rivers. Water Research, 32(11), 3245-3260. Hohenblum, P., Scharf, S., Gans, O., Moche, W. & Lorbeer, G. (2004). Monitoring of selected estrogenic hormones and industrial chemicals in ground waters and surface waters in Austria. Science of the Total Environment, 333, 185-193. Balcioglu, A. & Ötker, M. (2003). Treatment of pharmaceutical wastewater containing antibiotics by O3 and O3/H2O2 processes. Chemosphere, Vol. 50, Issue 1, 85-95. Paumann, R. & Vetter, S. (2003). Hormonwirksame Stoffe in Österreichs Gewässern – Ein Risiko? – ARCEM-Endbericht, Umweltbundesamt GmbH. Vienna. ISBN 3-85457695-1. (Endocrine disrupters in Austria‘s waters – a risk? – Austrian research cooperation on endocrine modulators) Federal Republic of Austria: QZV Chemie OG, Qualitaetszielverordnung Chemie Oberflaechengewaesser – (BGBl. II Nr. 96), 2006 (Quality Target Ordinance, Chemistry surface waters) European Community: Directive 2000/60/EC of the European Parliament and of the Council establishing a framework for Community action in the field of water policy, 23.10.2000 Scharf, S., Gans, O. & Sattelberger, R. (2002). Arzneimittelwirkstoffe im Zu- und Ablauf von Kläranlagen; BE-201, ISBN 3-85457-624-2Umweltbundesamt GmbH: Vienna. (Pharmaceutical substances in inflow and effluent of STP´s, Data material, Austrian federal environmental agency). DIN 38406-3 (2002). German standard methods for the examination of water, waste water and sludge – Cations (group E) - Part 3: Determination of calcium and magnesium, complexometric tritation, E3, German Institute for Standardization. Menapace, H. M., Diaz, N. & Weiß, S. (2008). Electrochemical treatment of pharmaceutical wastewater by combining anodic oxidation with ozonation. Journal of Environmental Science and Health, Part A, 43:8, 961-968. Umweltbundesamt: Parameterinformationsblatt Atrazin und Desethylatracin. Datenband Porengrundwasser, Vienna, 2006. Breckenridge, C. B., Werner, C., Stevens, J. T. & Sumner, D. D. (2008). Hazard Assessment for Selected Symmetrical and Asymmetrical Triazine Herbicides. In The Triazine Herbicides, 387-398.

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[18] Eldridge, J. C. & Wetzel, L. T. (2008). Mode of Action of Atrazine for Mammary Tumor Formation in the Female Sprague-Dawley Rat. In The Triazine Herbicides, 399411 [19] Snedeker, S. & Heather, C. (1999). Critical Evaluation of Atrazine´s Breast Cancer Risk. Program on Breast Cancer and Environmental Risk Factors in New York State. Cornell University, Ithaca. [20] Spark, K. M. & Swift, R. S. (2002). Effect of soil composition and dissolved organic matter on pesticide sorption. Science of the Total Environment, 298, 1-3, Amsterdam. [21] Austrian Drinking Water Ordinance, BGBl 304/2001, released on 21.08.2001 [22] Pupunat, L. & Rychen, Ph. (2002). Inactivation of Legionella with the DiaCell® Water Treatment Technology, Swiss Center for Electronics and Microtechnology Inc, CSEM Scientific Report. [23] Furuta, T., Rychen, Ph., Tanaka, H., Pupunat, L., Haenni, W. & Nishiki. Y. (2005). Application of Diamond Electrodes for Water Disinfection. In Diamond Electrochemistry, A. Fujishima, et al, Ed.; BKC Inc., Tokyo et Elsevier B.V., Amsterdam, 525-542. [24] Maier, D. (2008). Untersuchung zur Wirkung des von der Fa. pro auqa Diamantelektroden Produktion GmbH entwickelten AOP-Verfahrens bei der Reinigung von Karlsruher Abwasser. Karlsruhe: DVGW.

In: Fluid Waste Disposal Editor: Kay W. Canton, pp. 235-253

ISBN: 978-1-60741-915-0 © 2010 Nova Science Publishers, Inc.

Chapter 10

UTILIZATION OF BIOSOLIDS AS FERTILIZATION AGENTS ON AGRICULTURAL LAND: DO THE OBVIOUS BENEFITS OF RECYCLING ORGANIC MATTER AND NUTRIENTS OUTWEIGH THE POTENTIAL RISKS? Veronica Arthurson* Department of Microbiology, Swedish University of Agricultural Sciences, Box 7025, 750 07, Uppsala, Sweden.

ABSTRACT Treatment of wastewater, commonly performed at municipal sewage plants, generates sanitized water and sewage sludge. Anaerobic degradation of sewage sludge results in the production of different gases, including the economically valuable methane, and digested residue (biosolids) with potential value as a crop fertilizer. Traditionally, digested sewage sludge is disposed either into water, onto or into the earth or into the air. However, alternative exploitation of digested sewage sludge in agriculture has several advantages over commercial fertilizers, including environmental aspects benefiting agricultural sustainability and increased crop yield. Additionally, residue utilization is nearly always a cheaper option than disposal costs. Biosolids obtained from the treatment of municipal sewage sludge consist of a mixture of organic and mineral compounds that significantly affect soil microbial communities and their biogeochemical activities when applied as a crop fertilizer. The microorganisms influence soil quality through nutrient cycling, decomposition of organic matter and maintenance of soil structure, in turn, affecting agricultural and environmental quality, and subsequently, plant and animal health. Moreover, both soil and residue normally contain considerable quantities of microorganisms, including both beneficial and potentially human pathogenic species that may be supported by the new conditions in *

Corresponding author: . Email: [email protected]., Phone: +46 - 18 - 67 32 12. Fax: +46 - 18 67 33 92

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Veronica Arthurson the soil. Thus, soil amended with biosolids may present a modified microbial community composition after some time and, hence, a modified ecosystem function. At the end of the present chapter, we discuss whether the potential risks of recycling biosolids to agricultural cropland are acceptable for consumers, producers and scientific expertise, in view of the resulting alterations in soil microbial diversity, activity and accompanying functions. Furthermore, optimal ways of managing the recycling process to achieve the most favourable balance of benefits and risks for the community are highlighted.

ANAEROBIC DEGRADATION OF SEWAGE SLUDGE Wastewater refers to water that is anthropogenically affected in quality, and is divided into three major groups, specifically, domestic, industrial and storm runoff. Sewage, on the other hand, is the part of wastewater polluted by feces or urine, commonly derived from the municipal sewer systems. The concentration of decomposable organic materials is reduced in sewage via a series of treatments. Treatment of sewage at conventional municipal wastewater treatment plants usually involves primary settling in which heavy solids and associated colloidal wastes from clarifiers are removed as primary sludge, followed by secondary biological aeration that generates reclaimed municipal wastewater and secondary sludge, the latter mostly constituting waste activated sludge, i.e., a mixture of primary treated sludge and biological organisms. The insoluble solid residue obtained is termed ‗sewage sludge‘ and the primary and secondary sludge fractions collectively contain 60% to 80% organic matter (dry weight) constituting carbohydrates, fats and proteins. To further reduce the organic matter and water content of sewage solids, and minimize their putrescibility and concentrations of pathogenic microorganisms, anaerobic digestion is preferably performed (Table 1), whereby optimally 80% of the influent organic content is biologically degraded. The ideal input material for anaerobic digesters is a mixture of primary and secondary sludge, which provides optimal bacterial composition required for the degradation process. Apart from the bacteria required for digestion, other factors, such as pH of sludge, nutrient content and C:N ratio, are important parameters that require adjustment for achieving optimal degradation. For instance, a high C:N ratio would result in low methane production due to the acidic environment created by the presence of low levels of N. In general, the C:N ratio in sewage sludge is below 10, providing an acceptable pH, and consequently, minimal risk of inhibition of methane-forming bacteria based on acidic conditions. Nevertheless, methanogens need a certain amount of organic acids for meeting their nutrient requirements. To maintain maximum activity of the bacterial groups involved in degradation, including methanogens, and obtain high gas production rates, the pH range (optimally between 6.8 and 7.5), temperature, nutrient level, C:N ratio, and mixing should be strictly monitored throughout the process (Williams, 2005). The initial stage of anaerobic digestion involves the hydrolysis and solubilization of solids resulting in the conversion of carbohydrates, proteins and lipids to simple soluble sugars, amino acids and fatty acids, respectively. This step is performed by specific hydrolytic bacteria producing exoenzymes capable of breaking the chemical bonds between organic units that make up insoluble particulate solids outside bacterial cells. Once solubilized, organic substrates are further degraded within cells. At the second stage, soluble compounds

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produced in the previous step are degraded by facultative anaerobic and strictly anaerobic bacteria via a number of fermentative processes. This acid-forming stage results in the production of carbon dioxide, hydrogen gas, alcohols, organic acids (acetate), some organicnitrogen compounds and a few organic-sulfur compounds (Gerardi, 2003). Acetate, either produced via fermentation of soluble organic compounds or acetogenesis (acids and alcohols are degraded to acetate), subsequently serves as the main substrate for methane-forming bacteria, finally resulting in methane gas and carbon dioxide, i.e., biogas (Gerardi, 2003). Additional gases present at low concentrations in biogas include hydrogen, hydrogen sulphide, nitrogen and trace levels of hydrocarbons (Williams, 2005). Biogas has a calorific value of between 20 and 25 MJ/m3 (IEA Bioenergy, 1996), and can either be directly combusted to produce heat for the digestion process or used in power generation. In the latter case, the gas needs to be cleaned up to remove corrosive trace gases, moisture and vapors from the gas stream (Williams, 2005). In addition to the benefits of anaerobic degradation in terms of renewable energy production and reduction in sewage sludge volume, weight and putrescibility, the resulting residue (biosolids) may have potential utility as an agricultural fertilizer. The insoluble solid residue derived from sewage sludge, termed biosolids, generally consists of organic compounds, macronutrients, micronutrients, non-essential trace metals, organic micro pollutants and microorganisms (Singh and Agrawal, 2008, Kulling et al., 2001). Macronutrients in biosolids provide a source of plant nutrients, and the organic content confers beneficial soil conditioning properties (Logan and Harrison, 1995), emphasizing its potential value as a crop fertilizer.

DISPOSAL AND/OR UTILIZATION OF BIOSOLIDS A major current problem concerns management of the increasing amounts of organic waste in an environmentally acceptable manner. Traditionally, sewage sludge was dumped into the ocean, which is now regulated and prohibited by American and European legislations (USEPA 40CFR 503 and 86/278 EEC, respectively), resulting in fewer waste management options, including incineration and land application (e.g., as soil fertilizer or conditioner). Incineration poses several drawbacks from the environmental viewpoint due to high energy demand (in view of the need for dewatering), requirement of flue gas cleaning to avoid atmospheric gaseous emissions and a resulting ash product that needs to be treated and disposed. The use of municipal solid landfill sites for sludge disposal was prohibited in 2005 in EU countries, and will be phased out in a stepwise manner, resulting in lower levels of landfill gas (mainly methane and carbon dioxide derived from degradation of the organic waste fraction) leached to the atmosphere, and consequently, decreased global warming. Moreover, the leachate produced from biodegradation of landfill waste confers a potential risk of watercourse pollution. Hence, land application remains the most promising alternative for handling biosolids as a way of recycling the nutrients of organic waste from human activities back to the production and supply chain. However, biosolids need to be rigorously assessed for quality due to the high content of metals (cadmium, arsenic, copper, lead, mercury and zinc), persistent organic pollutants (the organochlorines aldrin, dieldrin, heptachlor, dichlorodiphenyltrichloroethane and lindane) and pathogenic microorganisms (bacteria, viruses, protozoa and helminths) to ensure that harmful elements are not transmitted

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to humans through entry into the food chain via crops or grazing animals. Sludge stabilization by anaerobic or even aerobic digestion usually results in Class B biosolids, which have limited immediate reuse potential in land application owing to the high content of human pathogens. On the other hand, class A biosolids are generated as a result of more stringent treatment, and therefore contain extremely low or non-detectable amounts of microbial pathogens. Consequently, sewage sludge intended for anaerobic digestion and subsequent land application requires further treatment (prior to or after the degradation procedure) to obtain an environmentally acceptable end product of Class A quality. The most well established ways to achieve this goal are lime treatment, composting, and/or heat treatment, of which the latter often is the most efficient, resulting in the elimination of most microbial pathogens. However, bacterial endospores present in digested sewage sludge (Bacillus spp. and Clostridium spp) are not destroyed by standard heat procedures, and require a highly expensive method involving at least two separate rounds of heating to ≥ 70°C for ≥ 1 h to ensure elimination. Primary heating activates the spores into vegetative forms, which eventually start to metabolize and replicate. The secondary heating step should kill these heatlabile bacteria, provided the incubation period between the two heating steps is sufficiently short to prevent the formation of new endospores. Alternatively, irradiation technology should be considered to meet the requirement for Class A biosolids, and further assessed to determine whether it presents an effective option to the above sanitization techniques. However, most endospore-forming bacterial species are indigenously present in soil, and thus the issue of whether application of heat-treated biosolids actually imposes an increased risk is debatable. In conclusion, if biosolids are to be applied to agricultural land as a Class B waste product, site and crop restrictions are necessary. On the other hand, if the residue is further subjected to one of the above sanitization techniques (e.g., lime treatment, composting or heat treatment/irradiation), it would be a safer product in terms of risk of disease transmission via microbial agents.

EFFECTS OF BIOSOLIDS ON SOIL FERTILITY AND MICROBIAL ACTIVITIES Biosolids are applied to agricultural fields as a soil amendment to improve the chemical and physical properties and facilitate recycling of valuable components, such as organic matter, N, P and other plant nutrients (Martinez et al., 2002). The need for mineral fertilizers may be eliminated with the use of biosolids as fertilizer for arable land (Sommers, 1977). The method of biosolid spreading on fields depends on the percentage of constituent solids versus water, which determines its final character as liquid or solid matter. In the case of liquid residue, the product is preferably applied by injection to a soil depth of 6-9 inches, resulting in reduced bioaerosols, odors, and minimized risk of runoff to surface waters (Maier et al., 2009). Alternatively, liquid and solid residue can be applied to the soil surface, followed by tilling into the soil (Maier et al., 2009). Mantovi et al. (2005) evaluated the use of anaerobically degraded sewage sludge on agricultural land, highlighting the long-term effects on soil fertility and crop yield. The group showed that biosolids applied at a normal rate resulted in crop yields similar to those obtained with mineral fertilizers. Upon use of higher rates of mineral dressing as well as biosolids, the

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excess supply of available N resulted in poor quality crops. The authors concluded that for best yield and crop quality, it is important not to exceed a rate of 5 Mg dry matter ha-1 yr-1 of liquid or dewatered sludge. In contrast, composted sewage sludge did not cause negative effects on crop quality when applied at higher rates. Moreover, twelve years of continuous landspreading of biosolids resulted in improved soil fertility (organic matter and nitrogen increase), but a small buildup of heavy metals (Cu and Zn). The study by Mantovi and colleagues supports the significant potential of treated sewage sludge as a crop fertilizer, although the possible negative effects of organic pollutants and heavy metal content in sewage sludge intended for agricultural land application need to be considered. Moreover, plants differ in their abilities to absorb sludge-borne metals from soil (Singh and Agrawal, 2008), which is further influenced by soil properties, including pH, redox potential, sesquioxide content, organic matter, and application rate of sludge (Hue and Ranjith, 1994). These parameters should be carefully evaluated and monitored, prior to the application of biosolids to arable land. In another field trial, Odlare et al. (2008) studied the changes in soil chemical and microbiological properties over a period of 4 years, following treatment of soil with different organic waste products. Very few significant differences in chemical and microbiological features were detected with several fertilizers of different origins. However, the authors showed that biogas residue, generated from co-digestion of source-separated household waste and food residues from restaurants and kitchens, augmented the microbial biomass (analyzed on the basis of substrate-induced respiration), proportion of active microorganisms, nitrogen mineralization, and potential ammonia oxidation. Soil treated with anaerobically digested sewage sludge displayed similar trends of increased microbial biomass, larger proportion of growing microorganisms and higher degree of nitrogen mineralization, compared to unfertilized control soil, although the data were not statistically significant throughout the treatments. Additionally, the study by Odlare et al. (2008) disclosed no negative effects of organic waste on the soil microbial parameters, and equivalent or better fertilization effects of all waste products, such as cow manure, pig slurry and mineral fertilizer.. Consistent with these trends, Banerjee et al. (1997) reported that soil amendment by sewage sludge augmented microbial, respiratory, and enzyme activities, and higher microbial biomass in sludge-amended soil is frequently observed (Parat et al., 2005, Dar, 1996). Increase in microbial biomass when soil microbiota are supplied with C, N and nutrients (i.e. biosolids) is expected, since most soil microorganisms are heterotrophs that depend on organic carbon for growth and energy. On the other hand, a decrease in biomass is likely if the biosolids applied contain toxic chemicals, such as organic pollutants. Sullivan and co-workers (2006b) observed no changes in total microbial biomass in response to biosolid application, which was explained as a rapidly acclimatizing soil microbial community whose overall biomass was sparsely affected by biosolids. Comparable results were reported by García-Gil et al. (2004). The group of Epstein (1998) studied the effects of sewage sludge soil application on water retention, hydraulic conductivity and aggregate stability. In their experiments, both raw and digested sludge increased the total soil water retention capacity, as well as soil hydraulic conductivity after 27 days of incubation. Epstein (1998) additionally reported an average of 34% stable aggregates after 175 days of incubation in sludge-treated soil, compared to 17% in untreated control soil, confirming the beneficial characteristics conferred by sewage sludge. Long-term application of biosolids evidently improves the physical features of soil, including

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saturated and unsaturated hydraulic conductivity, water retention capacity, bulk density, soil resistance to penetration, total porosity, pore size distribution, aggregation, and aggregate instability (Tsadilas et al., 2005, Aggelides and Landra, 2000). Sewage sludge commonly contains high amounts of human pathogenic bacteria excreted in feces and urine (Dudley et al., 1980, Larsen, 1995, Strauch, 1991). The health risks related to these pathogens in biosolids spread on agricultural land depend on prior sludge treatments applied, as well as their ability to maintain virulence during storage and field distribution. Biosolids subjected to anaerobic digestion and an additional heating/pasteurization step ( ≥ 70°C for ≥ 1 h; Class A biosolids) should not contain human pathogenic microorganisms. However, harmful gut microorganisms frequently remain in class B biosolids. A number of studies have focused on evaluating the survival of indicator organisms and pathogens introduced into soil via biosolids. In particular, limited survival of enteric organisms, such as Escherichia coli, in biosolid-amended arable soil has been documented in laboratory and field trials (Lang and Smith, 2007). For instance, Lang et al. (2007) investigated the potential survival of different enteric microorganisms in agricultural soil amended with biosolids, and reported that dewatered mesophilic anaerobically digested biosolids increased the level of E. coli in soil. However, the introduced E. coli declined rapidly, and their survival was limited to three months, irrespective of the environment or timing of sludge application. The authors concluded that pathogens applied to soil via biosolids decline to background values well within the cropping and harvesting restrictions imposed when sewage sludge is spread on farmland. According to traditional Swedish farming practice, anaerobically digested sewage sludge is spread on agricultural land prior to ploughing in late autumn, which would allow the three months of potential pathogen survival to pass before harvesting of plants, consistent with the conclusions of Lang and colleagues (2007). However, the use of sewage sludge (treated or untreated) on agricultural land in Sweden has been very restricted for some years now due to the reluctance of the food industry and, hence, the farmers to use it (www.lrf.se). To improve the quality of the sewage sludge, a Swedish project called ReVAQ, was recently initiated and the results are evaluated by Urban Water AB (Malmqvist et al., 2006). Instead of using the sludge as a fertilizer on arable land, de-watered digestion residues derived from sewage sludge are frequently employed as inorganic matter for land reclamation and/or cover material at closed landfill sites (Eriksen, 2009). Anaerobic digestion residue obtained from the degradation of source-separated household waste is frequently used as a biofertilizer in Sweden (www.avfallsverige.se), as the risk for spreading heavy metals is low. Table 1. Advantages and disadvantages of anaerobic versus aerobic degradation. Anaerobic digestion

Aerobic digestion

No oxygen requirement

Requires oxygen

Energy-rich methane (e.g., biogas) produced

Energy mainly released as heat

Slow (in comparison with aerobic degr.)

Fast (in comparison with anaerobic degr.)

More sensitive to upsets by toxic agents

Less sensitive to toxic agents

Produces smaller amounts of sludge due to lower energy yields of anaerobic bacteria

Larger mass of solids due to higher energy yields of aerobic cell metabolism

Enhanced degradation of xenobiotics and recalcitrant natural compounds Process start-up requires long periods of time

Modified from Maier et al. (2009).

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In a 20-year field study at the University of Arizona (Zerzghi, 2008, Pepper et al., 2008), the influence of land application of biosolids on the soil microbial community was evaluated by assessing microbial number, activity and diversity. Class B biosolids had no adverse affects on soil microbial number (Pepper et al., 2008), but increased microbial diversity (evaluated through cloning and sequencing of bacterial 16S rRNA genes) and activity (assessed from common microbial transformations, nitrification and sulfur oxidation). Moreover, no known pathogens were present in soil sampled 9 months after the last application of biosolids, and after 20 years of biosolid treatment, fecal coliform and total coliform counts did not exceed 3 MPN g-1 in the amended plots (Pepper et al., 2008). Lawlor and co-workers (2000) reported that soil amendment by biosolids affected overall diversity and biomass only to a limited extent, but had varying effects on individual species, functions and specific microbial parameters. These results were consistent with a study by Dennis and Fresquez (1989) reporting shifts in the microbial community structure as a result of biosolid application. Moreover, Sullivan et al. (2006a) showed that biosolids negatively affected the presence of arbuscular mycorrhizal fungi (AMF; measured using specific biomarkers), but augmented the relative abundance of both Gram-positive and Gram-negative bacteria. The decrease in AMF biomarkers and increase in bacterial markers may be explained by the concept that soil microbial communities of infertile ecosystems are frequently dominated by fungi, whereas those of more fertile, productive ecosystems (such as biosolids) predominantly contain bacteria (Kourtev et al., 2003, Grayston et al., 2004, Wardle et al., 2004). In addition, Sullivan and colleagues (2006b) showed that biosolid-amended soil microbial communities were able to utilize Biolog EcoPlate substrate more quickly (via higher mineralization activity), compared to microbial communities from control soils. Based on the results, the authors suggest that the metabolism of at least parts of the soil microbial community remains elevated for extensive periods after land application of biosolids. Evaluation of soil chemical properties at the end of the 20-year land application study in Arizona (Pepper et al., 2008, Zerzghi, 2008) revealed an increase in total and available soil phosphate concentrations, consistent with data obtained from other similar studies (Mantovi et al., 2005, Brendecke et al., 1993). In addition, total N was increased in soil amended with biosolids, and nitrate levels in both biosolid and fertilizer-treated soil exceeded 10 ppm NO3-N at most soil depths down to 150 cm, indicating a potential risk for nitrate pollution of ground water, irrespective of the use of biosolids or mineral fertilizers (Pepper et al., 2008). Hattori (1988) examined the potential relationship between sewage sludge decomposition and microbial numbers and activities in soil. The mineralization rate of organic C and N rapidly increased bacterial numbers and proteinase activity in soil to maximum levels within the first 3 days, followed by a rapid decline thereafter. Bact erial numbers were significantly correlated with enzymatic activities and mineralization rates of organic C and N, indicating that proteinase-producing bacteria certainly contribute to rapid degradation of proteins in sludge immediately after soil application (Hattori, 1988). These types of studies would greatly benefit from the development of new molecular approaches (see review by Arthurson, 2008) providing additional information on bacterial community structures, metabolic activities and responses to C and N mineralization rates, and ultimately, correlation of these parameters with the effectiveness and quality of digested sewage sludge as a crop fertilizer. Molecular methods also have a considerable

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advantage over traditional assays in terms of independence from cultivation of bacteria of interest, resulting in an improved representative picture of a community, including the ―viable but non-culturable‖ fraction (Maraha et al., 2004). Consequently, fast and reliable detection/identification tools require further development to ensure more thorough evaluation of the soil microbial communities and corresponding ecosystem functions affected by biosolids. In combination with long-term field evaluation of land application of biosolids, methodological developments should provide information on ways to benefit from the recycled plant nutrients in biosolids without disrupting microbial, physical and/or chemical properties crucial for the soil ecosystem.

ANTIBIOTIC RESISTANCE GENES A potential risk of applying and reusing organic matter and plant nutrients of biosolids in soil is the possible introduction of antibiotics and resistance genes to the soil and water environment. In a bacterial community confronted with an antibiotic, only a single bacterium with a genetic or mutational change conferring resistance is required for the cell to start proliferation, which would rapidly result in a large number of resistant bacterial cells. Consequently, a greater number of antibiotics present within the environment would result in higher likelihood of bacterial strains developing resistance. In a study performed by Stromberger and Coffin (2006), samples of aerobically and anaerobically digested biosolids contained three different classes of antibiotics and significant populations of oxytetracycline-resistant bacteria. Moreover, upon application of biosolids to agricultural soils cropped with winter wheat at high rates (5 tons acre -1) and incorporation into soil (20 cm depth), the oxytetracycline-resistant bacterial fraction was greater in biosolid-amended soil, compared to non-amended plots. A higher proportion of culturable soil bacteria (60%) was resistant to 1 ppm oxytetracycline in th e amended plots, compared to 27% in control plots, and 4% of soil bacteria were resistant to 10 ppm of the antibiotic in biosolid-amended plots, compared to less than 1% in nonamended plots (Stromberger and Coffin, 2006). Research by Stromberger and Coffin (2006) showed that biosolids contain antibiotics as well as antibiotic-resistant microorganisms. The ability of non-pathogenic bacterial strains to horizontally transfer their antibiotic resistance genes to human pathogenic strains in the environment or human hosts is another potential theory (Rensing et al., 2002, Dzidic and Bedekovic, 2003, Salyers et al., 2004). Hence, biosolids may constitute a carrying matrix for antibioticresistant bacteria transfer via food and/or water into the host (Pepper et al., 2008), and ultimately, resistance genes could be transferred to bacterial strains present in the host. However, in contrast to the above results (Stromberger and Coffin, 2006), Brooks et al. (2007) recently presented data from soil sampled before land application of biosolids, and up to 450 days following application, showing a negligible influence of biosolids on the occurrence of antibiotic resistance of bacteria in soil (Brooks et al., 2007, Pepper et al., 2008). Regardless to the extent to which biosolids are able to induce antibiotic resistance among bacterial strains in amended soil, it is vital to establish the occurrence, fate and effects of antibiotics throughout the ecosystem environment.

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CONCLUSION Do the Benefits of Recycling Biosolids Outweigh the Risks? The value of biosolids as a fertilizer has been recognized for decades. However, the related risks of transfer and persistence of human pathogenic microorganisms, disruption of soil biological processes, altered soil quality and/or fertility, as well as accumulation of persistent organic pollutants and heavy metals have not been investigated until relatively recently. Numerous studies have focused on the heavy metals and organic contaminants present in sewage sludge. However, limited information is available on risk analyses and strategies to monitor microbial activities in biosolid-amended soil. The human pathogenic microorganisms present in biosolids naturally vary depending on several factors, and reflect the incidence of specific diseases in the community where the sewage sludge are derived. Hence, the environmental risks of land application should be considered for each individual biosolid product, depending on the specific compositions. In this chapter, the microbial, physical and chemical properties of biosolid-amended soil is discussed, focusing on the potential alterations in these parameters following the major input of organic matter, plant nutrients and toxic elements, potentially present in the biosolids). The majority of studies performed to date reveal no or limited negative effects on soil microbial/physical/chemical parameters, and in many cases, show that biosolids exert equivalent or better fertilization effects, compared to mineral alternatives. As long as the crucial parameters (heavy metals, organic pollutants, human pathogenic microorganisms) are carefully monitored in each batch of biosolids and values are within the regulated limitations, at least Class A biosolids can be further exploited as a suitable crop fertilizer. However, thorough analyses are required to establish the effects of biosolids applied to soil at the microbial/enzymatic level. The majority of the publications reviewed in this chapter support an increase in soil enzymatic processes, such as nitrification, sulphur oxidation and dehydrogenase activity, as a result of biosolid amendment (Pepper et al., 2008, Zerzghi, 2008). Moreover, land application of biosolids have been shown to positively affect soil microbial communities, leading to increased soil microbial number, activity and diversity (Pepper et al., 2008, Zerzghi, 2008). Treatment of land with biosolids (and organic matter in general) improves soil infiltration, water holding capacity, total porosity and soil aggregation, and enhances the total organic carbon and nitrogen content (Epstein, 1998, Tsadilas et al., 2005, Aggelides and Landra, 2000, Zerzghi, 2008). In conclusion, land application of sewage sludge raises certain crucial environmental issues, such as accumulation of organic pollutants and heavy metals, persistence of human pathogens, leakage of excessive nutrients, and viral contamination of surfaces and groundwater. Regulation of these negative parameters by strict monitoring and further assessment, in conjunction with the several positive effects discussed above, may allow exploitation of biosolids as a beneficial fertilizer from an agricultural/environmental point of view. Injudicious amendment can have a potentially toxic impact, while application of biosolids with adjusted parameters should provide essential nutrients for agricultural crop growth, and may be a good environmental and economic alternative to mineral fertilizers. To minimize risks and optimize benefits from recycling biosolids, the entire recycling process must be carefully managed in an appropriate manner. For instance, accurate records

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of biosolid application should be maintained to enable retrospective investigations to establish the operating conditions in case of problems. Moreover, the limits of toxic elements that suppress soil microbial processes should be defined and monitored in the field, and rapid and reliable techniques identified, enabling monitoring and supervision of the important processes in biosolid-amended soil (Davis et al., 1986). Additionally, optimized methods should be used for detection of specific microbial groups commonly present in biosolids, with the potential to stimulate/inhibite critical key-functions in the soil ecosystem as the specific biosolids are applied to arable land. Optimization of current monitoring techniques, regulations and recommendations should significantly benefit the recycling of biosolid products, identified to be harmless, back to cropland, with the ultimate aim of implementation in wastewater treatment facilities and biogas plants worldwide.

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Odlare, M., Pell, M. & Svensson, K. (2008). Changes in soil chemical and microbiological properties during 4 years of application of various organic residues. Waste Manag, 28, 1246-53. Parat, C., Chaussod, R., Leveque, J. & Andreux, F. (2005). Long-term effects of metalcontaining farmyard manure and sewage sludge on soil organic matter in a fluvisol. Soil biol biochem, 37, 673-679. Pepper, I. L., Zerzghi, H., Brooks, J. P. & Gerba, C. P. (2008). Sustainability of land application of class B biosolids. J environ qual, 37, 58-67. Rensing, C., Newby, D. T. & Pepper, I. L. (2002). The role of selective pressure and selfish DNA in horizontal gene transfer and soil microbial community adaptation. Soil biol biochem, 34, 285-296. Sahlström, L. (2003). A review of survival of pathogenic bacteria in organic waste used in biogas plants. Biores tech, 87, 161-66. Salyers, A. A., Gupta, A. & Wang, Y. (2004). Human intestinal bacteria as reservoirs for antibiotic resistance genes. Trends microbiol, 12, 412-416. Singh, R. P. & Agrawal, M. (2008). Potential benefits and risks of land application of sewage sludge. Waste Manag, 28, 347-58. Sommers, L. E. (1977). Chemical composition of sewage sludges and analysis of their potential use as fertilizers. J environ qual, 6, 225-232. Strauch, D. (1991). Survival of pathogenic micro-organisms and parasites in excreta, manure and sewage sludge. Rev sci tech off int epiz, 10, 813-46. Stromberger, M. E. & Coffin, K. (2006). Impacts of biosolids-borne antibiotics on soil microbial communities. Annual meeting of the soil science society of America. Indianapolis, IN, November 12-16. Sullivan, T. S., Stromberger, M. E. & Paschke, M. W. (2006a). Parallel shifts in plant and soil microbial communities in response to biosolids in a semi-arid grassland. Soil biol biochem, 38, 449-459. Sullivan, T. S., Stromberger, M. E., Paschke, M. W. & Ippolito, J. A. (2006b). Long-term impacts of infrequent biosolids applications on chemical and microbial properties of a semi-arid rangeland soil. Biol fertil soils, 42, 258-266. Tsadilas, C. D., Mitsios, I. K. & Golia, E. (2005). Influence of biosolids application on some soil physical properties. Commun soil sci plant anal, 36, 709-716. Wardle, D. A., Bardgett, R. D., Klironomos, J. N., Setala, H., van der Putten, W. H. & Wall, D. H. (2004) Ecological linkages between aboveground and belowground biota. Science, 304, 1629-33. Williams, P. T. (2005). Waste treatment and disposal, John Wiley & Sons Ltd, West Sussex. Zerzghi, H. (2008). Sustainability of long-term land-application of Class B biosolids: Influence on soil microbial and chemical properties. Ph.D. diss. Univ of Arizona, Tucson, AZ.

In: Fluid Waste Disposal Editor: Kay W. Canton, pp. 249-269

ISBN: 978-1-60741-915-0 © 2010 Nova Science Publishers, Inc.

Chapter 11

INTEGRATED APPROACH FOR DOMESTIC WASTEWATER TREATMENT IN DECENTRALIZED SECTORS 1

Rani Devi1* and R. P. Dahiya1 Centre for Energy Studies, Indian Institute of Technology, Hauz Khas, Delhi, India-110016.

ABSTRACT The purpose of the present study was to design an integrated wastewater treatment system for a nalla (riverlet) flowing through Indian Institute of Technology Delhi (IITD), India, besides its cost estimation and comparison with the conventional wastewater treatment system. The design parameters for integrated aeration-cum-adsorption tank were worked out for 240 m3 / d flow rate of the wastewater. The important parameters used for the design included initial COD and BOD concentration in the influent, treatment time, adsorbent dose, pH, adsorbent particle size and the desired COD and BOD in the effluent after treatment as prescribed by Central Pollution Control Board, (CPCB) Delhi, India. All the design parameters of this system were similar to those of conventional system except for the replacement of aeration tank in conventional system by the aeration-cum-adsorption tank. The concentration of COD and BOD of the treated effluent by the integrated system were well within the permissible limits of CPCB standards (for COD it is 100 ppm and for BOD of 30 ppm) to discharge in the canal for irrigation purpose. It was worth mentioning here that the adsorbents used in the present study were based on discarded materials which were available free of cost. Of course, the cost of their transportation and processing should have been taken into account. The total cost estimated for the conventional system and the adsorption based system would be Rs. 198,312 and Rs. 141,275 respectively (including civil work, machinery, labour, adsorbent and miscellaneous). The cost difference for the two systems would be approximately Rs 57,037. This design of integrated system has resulted into saving of cost by 28 % over the conventional system. Thus, it is a good approach for saving of conventional energy in *

Corresponding author: Department of Energy and Environmental Sciences, Ch. Devi Lal University, Sirsa, Haryana, India. E-mail address: [email protected], Telephone no.: +91-9416987345.

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Rani Devi and R. P. Dahiya addition to saving the cost of treatment and can be applicable for any country for decentralized sector. Moreover, it is an open ended research and we can recommend more research by changing the adsorbents types and operating parameters to improve the model.

Keywords: Adsorption; Cost Estimation; Decentralized Sector; Wastewater Treatment.

1. INTRODUCTION Water is an important commodity for the survival of living species on earth. Quantity of water utilized has been an index for the quality of life of the people. More water utilized essentially leads to more wastewater generation, be it in the urban or rural settings. For improving the quality of life in decentralized sectors, the use of electricity, water supply, sanitation and other amenities should be raised significantly. It is well known that the wastewater of domestic origin typically contains organic matter, pathogens, suspended solids, nutrients (nitrogen and phosphorus) and other pollutants. For curtailing the environmental and health hazards, there is a need to bring these pollutants down to the permissible limits for its safe disposal (3, 6, 8, 12). Removal of the organic contaminants and pathogens from wastewater is of paramount importance (1, 5, 13). The conventional wastewater treatment technologies as adopted in industrialized nations are expensive to build, operate and maintain (4, 9, 10), especially for the decentralized communities. Usually there is an inadequate arrangement for handling this wastewater which either flowing in open channels or accumulates in low-lying areas or flows through natural open drains. Ranges of technologies are available for treating these wastewater pollutants. Research efforts are on (10, 14, 15) for the development of better treatment technologies suited to these decentralized communities. Keeping in view the requirements of suitability and money, it has become imperative to find less costly and easily adaptable treatment technologies. Adsorption based innovative technology (14, 15) proposal having low cost carbonaceous materials showed good potential, for COD and BOD removal from the domestic wastewater. The carbon content of adsorbents plays a significant role during the adsorption of organic impurities. The adsorption capacity increases with the increased carbon content of the adsorbent and such a trend has been observed by various investigators (7, 16). Studies have shown that the carbonaceous adsorbents prepared from discarded materials have potential for reducing the COD and BOD level (14, 15). Use of such adsorbents can either completely replace or supplement the aeration tank. The aim of this paper is to design an integrated wastewater treatment system for IIT Delhi nalla by incorporating conventional as well as adsorption based treatment system. Results of the adsorption experiments reported in our another paper are utilized in evaluating the design parameters of the adsorption reactor. We have also calculated and compared the costs of the conventional and integrated wastewater treatment systems for IIT Delhi nalla.

2. MATERIALS AND METHODS A part of the domestic wastewater discharged in open channels along the streets of an urbanized village named Katwaria Sarai, situated on the periphery of Indian Institute of

Integrated Approach for Domestic Wastewater Treatment in Decentralized Sectors 251 Technology (IIT), New Delhi, India, spills into a natural nalla (small drain). The drain flows through IIT campus, meets other open drains on the way and eventually joins the river Yamuna. For the present investigations, we have collected wastewater samples from five points across the nalla. The samples were stored at 2-3 0C to avoid any change in their physico-chemical characteristics. The important physico-chemical characteristics analyzed were pH, temperature, total solids (TS), total suspended solids (TSS), chemical oxygen demand (COD) and biochemical oxygen demand (BOD). The pH and temperature of the wastewater samples were measured on collection site. Total solids, total suspended solids, COD and BOD were analyzed in laboratory according to the methods prescribed by APHA handbook (2). The COD and BOD of the wastewater samples were measured in laboratory before and after treatment with the mixed adsorbent carbon prepared by mixing fly ash, saw dust, sugarcane bagasse, coconut coir and brick kiln ash in 1:1 ratio). Wastewater was treated under batch mode by using mixed adsorbent carbon for removing COD and BOD load from wastewater and the treatment conditions included; treatment time, adsorbent dose, ph, initial COD and BOD concentrations, agitation speed and adsorbents particle size.

2.1. Design Criteria for Conventional Wastewater Treatment Plant of IIT Delhi Nalla The parameters considered for designing a domestic wastewater treatment plant were, wastewater flow rate, physico-chemical characteristics of the wastewater, mass loading and desired characteristics of the effluent. From the wastewater discharge, flowrate was calculated and its average value was taken for designing the wastewater treatment plant for the nalla. The average design flow was the daily average wastewater flow rate per unit of wastewater generated from the source and it was 240 m3/day. The conventional wastewater treatment system for IIT Delhi nalla could broadly be divided in two units. The first is primary treatment unit including bar screen chamber, grit chamber and primary clarifier and second is secondary treatment unit including aeration tank, secondary clarifier and sludge handling bed. There is a transfer pump between the primary and secondary treatment units. The nomenclature used in the design equations were: volumetric flow rate as Q; volume, length, breadth, depth and wall thickness of the tanks as V, L, B, D and d respectively; retention time as t, wall surface area as A and volume of the building material as Vb with appropriate subscripts. The concept of common walls led to material saving. Two or more walls of the chambers could be shared with those of the bar screen chamber, grit chamber, primary clarifier, aeration tank, sludge drying bed. Bar Screen Chamber: Volume of the bar screen chamber (VBSC) was calculated by using equation 1, wastewater flow rate was taken 240 m3/day and retention time of ½ hr. VBSC = Q ts

1

For calculating the volume of material used, we assumed depth (Dsc) and breadth (Bsc) of the chamber as 1 m and 1.3 m respectively. The chamber length (Lsc) was then ~ 3.85 m along the flow direction. Number of walls (n) used in sharing were 2 and the volume of material used (Vmbsc) was calculate by using the equation numbering 2 as:

252

Rani Devi and R. P. Dahiya Vmbsc = n Lsc Dsc dsc + Lsc Bsc dfs

2

where dsc is the wall thickness with a value of 0.15 m and dfs is the foundation thickness with a value of 0.2 m for the bar screen chamber. Grit Removal Chamber: It share a common wall with bar screen chamber. Its bottom was downward sloping and away from the screen chamber so that the grit could settle at the end. It was cleaned manually. The important point in the design of the grit basin was that the flow velocity should neither be too slow nor too high. Flow velocity should be enough to scour out the settled organic matter and reintroduce it into the flow stream. Such a critical scouring velocity (vH) is given by the modified Shield‘s formula (Garg; 1998). For grit particles of 0.2 mm diameter and with sedimentation co-efficient of 0.986, the range of critical scour velocity was calculated from equation numbering 3 as: vH = 3 to 4.5

gdg (Ss 1)

3

where dg is the diameter of the grit particle, g is acceleration due to gravity and Ss is the sedimentation coefficient. vH thus calculated was 0.11 to 0.25 m/s. Like the bar screen chamber, the retention time in this section has been taken as ½ hr for an average flow rate of 240 m3/day and ¼ hr for peak flow rate of 480 m3/day. We have evaluated the grit chamber tank volume (Vgc) by using equation 1. Volume of building material (Vmg) was calculated using equation 2 and also the concept of one wall sharing was also maintained. The tank has a triangular sloping outlet and a slot. Primary Clarifier: Since the effluent generation varies throughout the day, it is necessary to hold the effluent in the primary clarifier for about 4 hours retention time to homogenize the quality and quantity of the effluent. We assumed the tank depth, sidewall thickness and thickness of foundation as 3 m, 0.15 m and 0.20 m respectively. We have calculated the volume of primary clarifier (VCS), surface area (Acs), sidewall (S) and volume of material used in primary clarifier (Vmcs) by using these assumed parameters. Transfer Pumps: Effluent collected in primary clarifier tank was transferred at a uniform flow rate to an aeration tank to avoid the shock loads to the system. Two pump sets were provided one functioning and other as standby having pumping capacity of 10 m3/hr and 4.5 m head with 1.5 hp rating. Aeration Tank: Effluent was transferred from primary clarification tank to the aeration tank at a rate of 10 m3/hr using transfer pump of 1.5 hp rating. Aeration tank is effective upto 85–90 % reduction of suspended solids, biological oxygen demand (BOD) and chemical oxygen demand (COD) from wastewater. In the activated sludge process we usually consider pH, BOD, BOD: N: P ratio and F/M ratio (this ratio is also known as the food to microorganisms ratio) as 7-7.5, 300 mg/l, 100:5:1 and 0.2:0.5 respectively. The volume of aeration tank (Va) with 240 m3/day flow rate of wastewater was calculated by using the equation given as number 4.

Integrated Approach for Domestic Wastewater Treatment in Decentralized Sectors 253 Va = Q (Bbod) / (Xfm) Y

4

By assuming depth of tank (D) as 3 m, area of the aeration tank (Aa), side of aeration tank (Sa) and BOD load (Bld) of tank were calculated by using equations numbering 5,6 and 7 respectively as: Aa = VA / D

5

Sa =

6

Aa

Bld = Q.B

7

The oxygen required (O2) for the aeration tank is 1.5 kg/kg of BOD load/day (Bld). So O2 required was calculated as: O2 required = 1.5 Bld

8

We assumed oxygen transfer rate (Otr) for fixed aerator was 1.35 kg/hp/hr. Thus horse power (H.P.) required for the aerator was calculated from equation numbering 9 as: H.P. = O2 required / day Otr

9

To calculate volume of material used in the aeration (Vma) tank, we have area of side wall (As)2 as (6.324)2 , depth of tank (D) was 3 m and area of foundation (Afa)2 as (6.47)2 m2 and we assumed thickness of foundation (dfa) as 0.2 m. So Vma was calculated using equation numbering 10 as: Vma = 3 [(Afa)2 – (As)2] + 0.2 (Afa)2

10

The amount of sludge required for recirculation was calculated with the help of equation numbering 11 as: QR = (QY)/(Xs-Y)

11

where QR is recirculated sludge (m3/d), Q is wastewater flow rate (m3/d), Y is mixed liquor suspended solid (MLSS) (mg/l) and Xs is activated sludge quantity from secondary clarifier. Secondary Clarifier: The mixed liquor passed through a sedimentation tank where separation of the activated sludge from aerated water took place. The settled activated sludge was removed from the clarifier by gravity and divided into two streams. One stream called return sludge was sent back to the aeration tank near the inlet. Here, it mixed with the incoming primary treated effluent and acted as seed for the formation of more activated sludge and simultaneously maintained the MLSS between 3000-3500 mg/l. The other stream was excess sludge and was sent to the digester for digestion or percolation off as manure or for using as land fill. The percolated liquor from the sludge was sent back to the process stream for treatment. The treated effluent might also be used for horticultural purposes. The

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area (ASC), volume (VSC) and retention time (t) of secondary clarifier tank was calculated by assuming depth (D) of tank 3.2 m by following equations numbering 12,13 and 14 as: ASC = (Q + QR) / Lr

12

VSC = ASC D

13

t = VSC / Q

14

Volume of material used in tank (Vmsc) was calculated by assuming thickness of foundation (dfsc) as 0.2 m by following equations numbering 15 and 16 as: / 4 Dsc = Vmsc = [

ASC

/ 4 (ASC D) +

= 4.04 m / 4 (Afsc dfsc)]

15 16

Sludge Drying Bed: Sludge formed in the clarifiers is dried in the sludge drying bed for its final disposal. For designing the sludge drying bed, we assumed breadth (Bdb) of 4 m, length (Ldb) of 4 m, depth (Ddb) of 1 m, volume of the bed (Vdb) as 16 m3 and wall thickness of the foundation (ddb) as 0.15 m and thickness of foundation (dfdb) 0.2 m. So the volume of material (Vmdb) used in the drying beds was calculated considering two wall sharing concept by following equation numbering 17 as: Vmdb = n ldb Ddb ddb + ldb Bdb dfdb

17

Total amount of material used in civil work From the above calculations, the total volume (VT) of material used in the civil work was calculated by following equation numbering 18 as: VT = Vmbsc + Vmg + Vmcs + Vma + Vmsc + Vmdb = 2.155 + 1.59 + 10.20 + 14.142 + 7.06 + 4.4

18

2.2. Design Parameters for Integrated Wastewater Treatment Plant for IIT Delhi Nalla We have worked out the designing of the integrated wastewater treatment plant for IIT Delhi nalla. Design of aeration tank was modified by incorporating adsorption process by adding discarded material based mixed adsorbent carbon and it was named as aeration-cumadsorption tank. Aeration-cum-adsorption tank was designed on the basis of results obtained from the adsorption experiments for COD and BOD reduction from domestic wastewater. All the units of the wastewater treatment system described in section 2.1 for IIT Delhi nalla will be similar except for the aeration tank which was replaced by aeration-cum-adsorption tank. Now secondary clarifier was not required and aeration-cum-adsorption tank took its place.

Integrated Approach for Domestic Wastewater Treatment in Decentralized Sectors 255 The design parameters for integrated aeration-cum-adsorption tank were worked out for 240 m3 / d flow rate of the wastewater. A vertical partition wall was raised to divide the aeration-cum-adsorption tank in two parts. A wire mesh filter was placed in the outlet slit of the tank to prevent the adsorbent loss and also some additional adsorbent added intermittently on daily basis to make up the lose. Horizontal partition was made in the middle of the tank which acts as a shelf. Discarded material based mixed adsorbent carbon was added on the horizontal shelves and also at the tank bottom in both sections. The wastewater coming from the primary settling tank was fed to the aeration-cum- adsorption tank from its top. A steady flow of air was bubbled through the liquid from the shelf and bottom of the tank. This agitated and homogenized the COD and BOD laden liquid and also contributed to their removal through the aeration process. The volume of the aeration-cum-adsorption tank calculated. The volume of the construction material used for the aeration-cum-adsorption (Vmaca) tank was calculated by assuming wall thickness (daca) and foundation thickness (dfaca) of the tank as 0.15 m and 0.2 m respectively. Total volume of the construction material used was calculated by following equation numbering 19 as: Vmaca = Laca Daca daca + Laca Baca dfaca

19

The total volume of the material used in constructing the aeration-cum-adsorption based wastewater treatment system included the material for civil work in all the units of the treatment system as given in section 2.1 and 2.2. There would be no secondary clarifier with the aeration-cum-adsorption unit but there were two units of aeration-cum-adsorption tank as described earlier. The total volume of the construction material (VTmaca) was given by following equation numbering 20 as: VTmaca = Vmbsc + Vmg + Vmcs + (2

Vmaca) + Vmdb

20

2.3. Cost Estimation for Adsorption Based Wastewater Treatment System The estimate was made of approximate cost for the construction of conventional as well as integrated wastewater treatment plant for a small community having 240 m3/day flow rate. The cost estimation for the adsorption based wastewater treatment system included the estimated cost of civil works, electricity, mechanical equipments, aeration equipment and the cost of adsorbent.

3. RESULTS AND DISCUSSION The important wastewater sampling sites and their physico-chemical characteristics in term of temperature, pH, total solids (TS), total suspended solids (TSS), chemical oxygen demand (COD) and biochemical oxygen demand (BOD) were shown in Table 1. Important parameters used for designing a domestic wastewater treatment plant for IIT Delhi Nalla were given in Table 2. Figure 1 showed the main units for the treatment of the wastewater discharged in IIT Delhi nalla and Figure 2 showed detailed diagram of units for this treatment

256

Rani Devi and R. P. Dahiya

plant. Figure 3 depicted the hydraulic flow rate for each unit of the conventional wastewater treatment plant for IIT Delhi nalla and Figure 4 showed the sectional drawing for each unit of the conventional wastewater treatment plant for IIT Delhi nalla. The concept of common walls led to saving in material. Thus, two or more walls sharing concept could be used for bar screen chamber, grit chamber, primary clarifier, aeration tank, secondary clarifier and sludge bed for saving the cost. Table 1. Wastewater characteristics of the sample collected from IIT Delhi nalla at different points. Sites

Sampling point

pH

Temp. (oC)

TS (mg/l)

TSS (mg/l)

COD (mg/l)

BOD (mg/l)

A

Wastewater collected from a point near the Micro Model Center

8.2

31.4

1140

650

980

475

B

Wastewater collected from combined stream near the bridge after Micro Model Center, nearly 100 m from point A

7.9

31.3

1115

570

995

478

C

Wastewater collected nearly 100 m from point B

7.4

31.0

1100

490

1005

490

D

Wastewater collected from combined stream near second bridge in old campus, nearly 50 m from point C

7.1

31.1

1050

450

1080

505

E

Wastewater collected from the outlet (static zone) near the IIT gate, 125 m from point D

7.5

31.3

1020

320

1080

505

Table 2. Important parameters used for designing a domestic wastewater treatment plant for IIT Delhi nalla. Parameters

Values 3

Actual flow rate (m /day)

200

Over design (%)

20 3

Average design flow rate (m /day)

240

Peak flow rate (m3/day)

480

Wastewater characteristics pH Total solids (mg/l) Total Suspended solid (mg/l) COD (mg/l) BOD (mg/l)

7.1 1140 650 1080 505

Integrated Approach for Domestic Wastewater Treatment in Decentralized Sectors 257

1: the Main units for the treatment of the wastewaterindischarged IIT Delhi nalla. Figure 1. MainFigure units for treatment of the wastewater discharged IIT Delhiinnalla.

Figure 2:

Detailed diagram of units for the treatment of the

Figure 2. Detailed diagram of units for the treatment of the wastewater discharged in IIT Delhi

Taking 240 m3/day flow rate (Q) and ½ hr retention time for this average flow in the chamber, the volume of bar screen chamber came out to be 5 m3 and material for constructing the tank walls with a foundation of 0.2 m depth yielded 2.155 m3. Like the bar screen chamber, the retention time in grit removal chamber had been taken as ½ hr for an average flow rate of 240 m3/day and ¼ hr for peak flow rate of 480 m3/day and volume of tank calculated was 5 m3 and total volume of the construction material used was 1.59 m3 when wall sharing concept was used. Volume of the primary clarifier for 4 hour retention time came out to be 80 m3 and volume of the material used was 10.20 m3. Two pump sets were provided using one as functioning and other as standby and both having pumping capacity of 10 m3/hr and 4.5 m head with 1.5 hp rating as shown in Figure 2.

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Rani Devi and R. P. Dahiya

Figure 3. Schematics of hydraulic flow rate for each unit of the conventional wastewater treatment plant for IIT Delhi nalla. Figure 3: Schematics of hydraulic flow rate for each unit of the conventional wastewater treatment plant for IIT Delhi nalla.

Figure 4. 4: Flow diagram to show drawing for each unit unit of the wastewater Figure Flow diagram to sectional show sectional drawing for each of conventional the conventional wastewater treatment plant forforIIT treatment plant IITDelhi Delhinalla. nalla.

Volume of aeration tank was calculated as 120 m3 and area of the aeration tank (Aa), side of tank (Sa) and BOD load (Bld) of tank were 40 m2, 6.324 m and 72 kg/day respectively. For the present design, value of oxygen required as calculated from equation 8 was 108 kg/day. And thus for the aeration tank power of 2.9 hp was required as calculated from equation 9. But to include stringent requirements, motor power was of 5 hp. Volume of material used in aeration tank came out to be 14.142 m3 as calculated from equation 10. Quantity of recirculated sludge as calculated by using equation 11 was 88 m3/day and the ratio for sludge

Integrated Approach for Domestic Wastewater Treatment in Decentralized Sectors 259 recirculated to wastewater flow rate would be 0.366. The volume of secondary clarifier tank was 52.48 m3 with retention time of 5.2 hours as calculated by using equation 12, 13 and 14 and volume of material used in secondary clarifier was 7.06 m3. Volume of the sludge drying bed as calculated by equation 17 was 4.4 m3. The overall design dimensions of conventional wastewater treatment plant for IIT Delhi nalla are shown in Table 3. The overall design of this plant was presented by Figure 5. The total volume (VT) of material used for conventional wastewater treatment for IIT Delhi Nalla as calculated for the civil work was 39.55 m3. The design parameters for integrated aeration-cum-adsorption tank were worked out for 240 m3 / d flow rate of the wastewater. The important parameters used for the design were presented in Table 4 and included initial COD and BOD concentration in the influent as 1080 mg/l and 505 mg/l respectively, treatment time of 4 hours, adsorbent dose of 1400 kg, pH 7, 0.053 mm and the desired COD and BOD in the effluent after adsorbent particle size of treatment as 100 mg/l and 30 mg/l as prescribed by Central Pollution Control Board (4), Delhi, India. The dimensions calculated for aeration-cum-adsorption tank were 3 m 4 m 3.5 m as shown in Table 5. The design of the aeration-cum-adsorption tank was shown in Figure 6. Volume of construction material used for aeration-cum-adsorption tank was 4.6 m3. The final COD and BOD concentration in the effluent after treatment were 44.6 mg/l and 10.26 mg/l respectively and that was well below the permissible limits of Central Pollution Control Board, Delhi, India. Table 3. Overall design parameters for conventional wastewater treatment plant for IIT Delhi nalla. Name

Size in m (L B D)

Bar screen chamber Grit Chamber Collection sump Aeration tank Conventional clarifier Sludge drying beds

3.85 x 1.3 x 1 3.31 x 1.3 x 1 5.16 x 5.16 x 3 6.32 x 6.32 x 3 4.6 x 4.6 x 3.2 4 x 4 x1

Volume (m3) 5 5 80 120 52.8 16

Retention time (hrs) 0.5 0.5 8 12 4 –

Table 4. Parameters of the aeration-cum-adsorption tank using the discarded material based mixed adsorbent carbon for IIT Delhi nalla. Sr. No. 1. 2. 3. 4. 5. 6. 7.

Operating parameters Treatment time (hr) Adsorbent quantity (kg) pH Initial COD and BOD conc. (mg/l) Agitation speed (rpm) Adsorbent particle size (mm) Desired COD and BOD conc. (mg/l)

Values 4 1400 7 1080/505 600 0.053 100/30

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Rani Devi and R. P. Dahiya

Figure 5.Figure Overall5:dimensions wastewater plant for IIT Delhi nalla. Overall of dimensions of treatment wastewater treatment plant for IIT Delhi nalla. Vertical partition

3m 1.5 m

Wastewater inlet

Final treated effluent to distribute Compressed air

Compressed air

Horizontal perforated Shelf

Adsorbents

1.5 m Compressed air

1.5 m

0.75 m

0.75 m

Figure 6. Dimensions of aeration-cum-adsorption tank.

1.5 m

Figure 6: Dimensions of aeration-cum-adsorption tank.

Integrated Approach for Domestic Wastewater Treatment in Decentralized Sectors 261 Table 5. Overall design parameters for integrated (adsorption-cum-aeration) wastewater treatment plant for IIT Delhi nalla. Name

Size in m (L B D)

Bar screen chamber Grit Chamber Primary Clarifier Aeration-cumadsorption tank Sludge drying beds

3.85 x 1.3 x 1

Volume (m3) 5

Retention time (hrs) 0.5

3.31 x 1.3 x 1 5.16 x 5.16 x 3 (3 4 3.5) x 2

5 80 42 x 2

0.5 8 4

4 x 4 x1

16

–

The cost estimation for the adsorption based wastewater treatment system described above was a complex task. It included the estimated cost of civil works, electricity and mechanical equipments besides aeration equipment and the cost of the adsorbent. The volume of material used for the civil work of the aeration-cum-adsorption based wastewater treatment system was 27.54 m3 and for the conventional system as given in section 5.3 was 39.55 m3. The approximate cost of construction material was taken as Rs. 3000 per m3. The cost of construction and civil works then came out to be Rs. 118,650 for the conventional system and Rs. 82,620 for the adsorption based system. Adding 25 % as additional cost including the labour cost and overhead charges etc., then the cost of construction and civil works came out to be Rs. 148,312 for the conventional system and Rs. 103,275 for the adsorption based system. The cost of the mechanical system and electrical motors / pumps for the complete system could be worked out from the specific choice of the pumps and aerator. It was, however, expected that the capital cost for these equipments, other than the aerator, would be approximately Rs. 25,000 for the conventional 240 m3 / d wastewater treatment system. The aerator together with its motor would cost additional amount of nearly Rs. 25,000. For the adsorption based unit, the aerator could work with a smaller compressor powered by 1 hp motor. Its cost would be about Rs. 10,000. It was worth mentioning here that the adsorbents used in the present study were based on discarded materials which were available free of cost. Of course, the cost of their transportation and processing should have been taken into account. For 30 days charge of adsorbents and its make up material an amount of Rs. 3000 might be required. The total cost estimated for the conventional system and the adsorption based system would, therefore, be Rs. 198,312 and Rs. 141,275 respectively. The cost difference for the two systems would be approximately Rs 57,037.

CONCLUSIONS It is concluded from this paper that the replacement of aeration tank of conventional system by aeration-cum-adsorption tank can be a good option for wastewater treatment for decentralized communities. Due to the use of discarded material based carbonaceous adsorbents, the COD and BOD concentration has been reduced considerably and these limits

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are well within permissible limits of CPCB. It has also resulted into reduction of cost by approximately 28 %. Hence the proposed design of domestic wastewater treatment plant can be adopted for decentralized communities of the world.

REFERENCES [1] [2] [3] [4] [5] [6] [7] [8] [9] [10] [11] [12]

[13] [14]

Ali, M. & Deo, N. (1992). Effect of pH on adsorption process of chromium (VI) with a new low cost adsorbent, Indian Journal of Environmental Protection, Vol. 12(3), 202209. American Public Health Association (APHA), (1989). Standard methods for the analysis of water and wastewater, 17th Edn., Washington, DC. New York. Andrew, B., Xiaodong, S., Robert, G. & Edyvean, J. (1997). Removal of coloured organic matter by adsorption onto low cost-waste material, Water Res., Vol. 31(8), 2084-2092. Central Pollution Control Board (CPCB) (2000). Water quality parameters of wastewater treated effluent for irrigation purpose. Chen, P. H. (1997). Adsorption of organic compounds in water using a synthetic adsorbent, Env. Int., Vol. 23(1), 63-73. Garg, S. K. & Garg, R. (1998). Sewage disposal and air pollution engineering, Khana Publications, Delhi. Lier, J. B. V. & Lettinga, G. (1999). Appropriate technologies for effective management of industrial and domestic waste waters: the decentralized approach, Water Sc. Technol., Vol. 40, Issue 7, 171-183. Manju, G. N., Raji, C. & Anirudhan, T. S. (1998). Evaluation of coconut husk carbon for the removal of arsenic from water, Water Res., Vol. 32(10), 3062-3070. Mazumder, S. C. B. & Kumar, K. (1999). Removal/recovery of acetic acid from wastewater by adsorption on bagasse and coconut jute carbon, Indian J. Env. Hlth, Vol. 41(3), 170-175. Mazumder, D. & Roy, B. (2000). Low cost options for treatment and reuse of municipal wastewater, Indian J. Env. Prot., Vol. 20(7), 529-532. Piet, N. L., Piet, M. V., Lode, S. & Willy, H. V. (1994). Direct treatment of domestic wastewater by percolation over peat, bark and woodchips, Water Res., Vol. 28(1), 17-26. Rani Devi, Dahiya, R. P. & Gadgil, K. (2002). Investigation of coconut coir carbon and sawdust based adsorbents for combined removal of COD and BOD from domestic waste water, Water and Environmental Management Series, Inter. Water Assoc., 12091218. Rani Devi & Dahiya, R. P. (2006). Chemical Oxygen Demand (COD) Reduction in Domestic Wastewater by Fly Ash and Brick Kiln Ash, J. Water, Air and Soil Poll., Vol. 174(1-4), 33-46. Rani Devi, Dahiya, R. P., Ashok Kumar & Vijender Singh (2007). Meeting energy requirement of wastewater treatment in rural sector, International Journal of Energy Policy, Vol. 35, Issue 7, 3891-3897.

Integrated Approach for Domestic Wastewater Treatment in Decentralized Sectors 263 [15] Rani Devi & Dahiya, R. P. (2008). COD and BOD removal from domestic wastewater generated in decentralized sectors, Bioresource Technology, Volume 99, Issue 2, 344349. [16] Metcalf Eddy (2000). Environmental Engineering. New Delhi.

In: Fluid Waste Disposal Editor: Kay W. Canton, pp. 265-284

ISBN: 978-1-60741-915-0 © 2010 Nova Science Publishers, Inc.

Chapter 12

BIODEGRADATION CHARACTERISTICS OF WASTEWATERS 1

Fatos Germirli Babuna*1 and Derin Orhon≠2 Istanbul Technical University, Faculty of Civil Engineering, Environmental Engineering Department, Istanbul, Turkey 2 Turkish Academy of Sciences, Ankara, Turkey

ABSTRACT The objective of this chapter is to put forward an overview of biodegradation characteristics of wastewaters by emphasizing the significance of COD fractionation. Recalcitrant COD fractions of effluents can be used as a tool to evaluate whether discharge standards can be met with a prescribed biological treatment. Moreover, the appropriate type of biological treatment applicable to the wastewater under investigation can be addressed and the performance of an existing biological treatment system can be appraised with reference to inert COD fractions. Besides recalcitrant COD fractions of segregated industrial effluent streams can be regarded as an essential input of a sound industrial wastewater management strategy adopting minimization at source philosophy. Last but not least, data on COD fractions can be used as a solid source of information for modelling studies that define the design and performance of biological treatment systems. In this context, COD fractionation data on a wide spectrum of activities ranging from various industrial sectors to hotels is presented. Segregated industrial wastewater streams together with domestic sewage and end-of-pipe industrial effluents are evaluated in terms of their biodegradation characteristics.

1. INTRODUCTION Research activities focused on the understanding of the biodegradation characteristics of wastewaters are among the significant developments shaping environmental sciences in the * e-mail: [email protected] ≠ e-mail: [email protected]

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last several decades. Characterizing the organic content of effluents via COD parameter has remarkable advantages, as biodegradable COD provides an electron and energy balance between organic matter, biomass and oxygen utilized. This requires reliable determination of biodegradable fraction. Within the existing legislative framework it is possible to meet the effluent discharge standards defined in terms of COD by applying conventional biological treatment. In this manner domestic sewage and industrial effluents are handled as if they have similar characteristics. The effluents directed towards a biological type of treatment are composed of basically biodegradable and non biodegradable fractions yielding a treated effluent COD level under the discharge standards. Since all the biodegradable COD fractions are removed within biological treatment, COD remaining in the treatment plant outlet only contains the recalcitrant organics initially present in the wastewater itself with an additional supplement of residual microbial products generated during the course of biochemical reactions that take place in biological treatment. It should be kept in mind that currently, compliance with discharge standards depends on an evaluation based on total COD values. Such a traditional approach can be tolerated for domestic effluents. However it may cause remarkable negative impacts when applied to industrial effluents that are likely to contain an array of chemicals with various biodegradation characteristics. Identification of COD fractions with different biodegradability characteristics has been a turning point in the appraisal of wastewater characteristics. In this context, biodegradable and recalcitrant COD fractions have been experimentally assessed. The mentioned COD fractions together with experimentally determined kinetic and stoichiometric coefficients have been incorporated as inputs into new models structured to define the design and performance of biological treatment systems i.e. activated sludge. From a different perspective, biodegradation characteristics of segregated industrial effluent streams should be considered as backbone information leading to a sound industrial management strategy. At least a part of the countless numbers of chemicals used during manufacturing processes is discharged with the waste streams. Therefore, chemicals having a recalcitrant nature are introduced into the environment as they by-pass biological treatment plants without being removed. A new management protocol adopting waste minimization at source philosophy must be developed for handling such industrial wastewater streams. For this purpose first of all, auxiliary industrial chemicals generating highly recalcitrant discharges can be substituted by biodegradable ones. If such an application is not possible due to case wise reasons, then a specific treatment, preferably not targeting a complete removal but instead increasing the biodegradable fraction can be prescribed for these non biodegradable segregated industrial streams. These recalcitrant wastewater streams should be mixed with others only after passing through the mentioned specific treatment. In this context this chapter intends to provide an overview of biodegradation characteristics of wastewaters by presenting data on both domestic sewage and industrial effluents.

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2. MEASUREMENT OF ORGANICS IN WASTEWATERS 2.1. Conventional Parameters Wastewaters either of domestic or industrial origin contain a large spectrum of various organic compounds with different oxidation states. The majorty of these organic compounds cannot be quantified or even identified individually on a routine basis. Therefore collective parameters namely, total organic carbon (TOC); biochemical oxygen demand (BOD); and chemical oxygen demand (COD); are traditionally recognized to characterize the overall organic content of effluents. The information these collective parameters provide differs from each other. A brief evaluation on these collective parameters is presented below. TOC can be considered as a direct and precise parameter for measuring the total organic content of wastewaters. However the usage of TOC has some drawbacks: TOC does not differentiate between biodegradable and non biodegradable organic compounds. This fact no longer represents a major disadvantage as a technique for measuring the total biologically degradable organic carbon has been developed (Grady et al., 1999). The most significant deficiency of the TOC test on the other hand is; unlike BOD or COD; it does not indicate the oxidation state of the organic matter. Furthermore TOC parameter requires sophisticated and costly analytical instrumentation. What is the importance of getting information on the oxidation state? The energy available in an organic compound is proportional to its oxidation state. Because of that a highly reduced compound will provide more energy as compared to an equal mass of a highly oxidized compound. In other words, a highly reduced compound has a higher pollution potential than that of a highly oxidized one. The oxidation state of a compound is defined by the number of electrons available for transfer. The number of electrons can be translated directly into the amount of oxygen required to oxidize the compound to carbon dioxide and water, with 8 grams of oxygen being required per electron. This mentioned quantity of oxygen is defined as the oxygen demand that is used as a tool to measure the pollution potential of a wastewater. BOD parameter depends on assessing the pollution potential of a wastewater sample via measuring the amount of oxygen utilized during the growth of aerobic organotrophic microorganisms on available organic carbon source. The usage of BOD parameter as a sole reliable source of information is questionable since the following disadvantages and restrictions are involved: BOD parameter indicates an unknown fraction of the total organic content due to acclimation problems related to the used seed. Furthermore because of the potential existence of inhibitors in the wastewaters lower BOD values than actual can be obtained. These two factors represent significant drawbacks especially for effluents of industrial origin. Besides the conditions of the BOD test bottle sustaining a static culture, low organic matter concentration, low microorganism concentration, diminishing concentrations of dissolved oxygen, etc (Green et al., 1981) are quite different from those in biological treatment plants. In short, adoption of BOD can cause misleading results as only an unknown portion of the organic carbon in a wastewater sample is measured by this parameter. COD is a reliable parameter for measuring the oxygen equivalent of organic matter that is prone to oxidation by a strong chemical oxidant. COD covers almost all organics serving as the carbon or the energy source in heterotrophic growth. The relative magnitude of a few

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aromatic compounds (i.e. toluene, pyridine, etc.) not oxidized in the COD test can be considered as insignificant and therefore neglected in most effluents. On the other hand, COD cannot differentiate biodegradable organics from the non biodegradable compounds that are also oxidized in the test. However this issue no longer poses a major disadvantage since separate experimental protocols are developed for the identification of biodegradable and non biodegradable COD fractions under both aerobic and anaerobic conditions. The details of these experimental protocols based on either observing batch reactors or respirometric techniques can be found in literature (Germirli et al., 1991 and 1993; Germirli Babuna et al., 1998a; Ince et al., 1998; Orhon et al., 1999a; Orhon and Okutman, 2003). COD is often preferred to BOD and TOC as it provides an electron and energy balance between the organic matter, biomass and oxygen utilized with the stipulation that its biodegradable fraction is determined. Alhough brief evaluation given above clearly recognizes COD parameter as the most trustworthy method for reflecting the organic content of effluents, the adoption of BOD5/COD ratio is still used as a preliminary index of biodegradability. It is stated in literature that the effluents with BOD5/COD ratios higher than 0.4 may be considered as entirely biodegradable (Chamarro et al., 2001). Table 1 outlines BOD5/COD ratios of selected industrial wastewaters together with segregated industrial discharge streams carrying some auxiliary chemicals and domestic sewage. Table 1. BOD5/COD Ratios related to different sources SOURCE Industrial Wastewater Confectionery Plant I Confectionery Plant II Confectionery Plant III Slaughterhouse Poultry Processing Integrated Dairy Processing Tannery* Yogurt&Buttermilk Processing Cotton Knit Fabric Plant I Cotton Knit Fabric Plant II Cotton Woven Fabric Corn Wet Mill Segregated Industrial Stream Containing Carrier I Carrier II Biocide I Biocide II Natural Tannin Synthetic Tannin Lignin sulphonate Naphtalanesulphonate Domestic Sewage

* plain settled wastewater ** due to negligible BOD5

BOD5/COD

Reference

0.65 0.79 0.74 0.44 0.59 0.50 0.42 0.66 0.17 0.47 0.55 0.74

Orhon et al., 1995 Orhon et al., 1995 Orhon et al., 1995 del Pozo et al., 2003 Eremektar et al., 1999a Andreottola et al., 2002 Orhon et al., 1999b Orhon et al., 1993 Orhon et al., 1992 Orhon et al., 1992 Orhon et al., 1992 Eremektar et al., 2002

0** 0** 0** 0** 0.08 0** 0.15 0.03 0.46

Arslan-Alaton, et.al., 2006a; 2007 Arslan-Alaton, et.al., 2006a; 2007 Arslan-Alaton, et.al., 2006a; b Arslan-Alaton, et.al., 2006a; b Koyunluoglu et al., 2006 Koyunluoglu et al., 2006 Germirli Babuna et al., 2007c Germirli Babuna et al., 2007b Orhon et al., 1997

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As evident from the table certain industrial effluents (i.e. one textile plant dealing with cotton knit fabric manufacturing operations) and almost all of the segregated industrial streams containing the listed auxiliaries have very low BOD5/COD ratios indicating their recalcitrant and/or inhibitory and/or toxic nature. An increase observed in BOD5/COD ratio of a wastewater after being subjected to a partial chemical oxidation is commonly attributed to increased biodegradation. In other words, the potential of a chemical oxidation process to improve the biodegradability characteristics of an effluent is often monitored and consequently appraised by observing the BOD5/COD ratio. However the above given discussion on BOD parameter clearly indicates the misleading nature of such an evaluation. Organics or recalcitrant organics or inhibitory substances or toxic compounds, totally or partly, might be removed by partial chemical oxidation. On the other hand partial chemical oxidation might generate toxic and/or inhibitory by-products as well (Sharma et al., 2007). The BOD5/COD ratio presenting the net effect of all these possible mechanisms has to be handled very carefully and used in a limited manner by considering its deficiencies. Therefore at this stage, it is recommended to enrich such preliminary rough evaluations by conducting COD fractionationation experiments that yield more credible information.

2.2. Cod Fractionation The COD parameter reflects the amount of total organics in effluents by establishing appropriate correlations among substrate, biomass and dissolved oxygen in terms of electron equivalance. More precise information can be obtained when organics are differentiated in terms of their biodegradation rates. In this context, the total COD, CT1, in wastewaters can be divided into of two main components: The total non biodegradable or inert COD, CI1 and the total biodegradable COD, CS1: CT1=CS1+CI1 The inert COD can further be categorized in two subgroups identifying soluble and particulate fractions as soluble inert COD, SI1 and particulate inert COD, XI1: CI1=SI1+XI1 The total biodegradable COD, CS1 similarly comprises three fractions conveniently differentiated as readily biodegradable COD, SS1, rapidly hydrolysable COD, SH1, and slowly hydrolysable COD, XS1: CS1=SS1+SH1+XS1 Figure 1 shows major COD fractions of effluents determined by means of experimental evaluation. Similarly, Figure 2 illustrates the COD fractions of wastewaters after passing through a biological treatment process.

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CT1 Total Influent COD

CS1 Total Biodegradable COD

SS1 Readily Biodegradable COD

SH1 Rapidly Hydrolyzable COD

CI1 Total Inert COD

XS1 Slowly Hydrolyzable COD

SI1 Soluble Inert COD

XI1 Particulate Inert COD

Figure 1. Major COD fractions in wastewaters

ST Total Soluble COD

SS+SH Soluble Biodegradable COD

SI1 Soluble Inert COD of Influent Origin

SP Soluble Inert Microbial Products

(a) Soluble COD fractions

XT Total Particulate COD

XH Viable Heterotrophic Biomass

XS Particulate Biodegradable COD

XI1 Particulate Inert COD of Influent Origin

XP Particulate Inert Microbial Products

(b) Particulate COD fractions Figure 2. Major soluble and particulate COD fractions in the effluent of biological treatment

The effluent of a properly designed and well operated biological treatment system does not contain any biodegradable COD fractions as all the biodegradable organics will be removed within the treatment plant. The particulate inert COD, XI1 and particulate inert microbial products, XP, are entrapped within the biological sludge and leave the biotreatment

Biodegradation Characteristics of Wastewaters

271

system by sludge wasteage. In this respect among other COD fractions, the soluble inert COD gains importance since it by-passes the biological treatment system without being involved in the biochemical reactions. Soluble residual (inert) microbial products, SP, together with the inert COD of influent origin, SI1, jointly control the magnitude of effluent soluble COD. Therefore, the effluent soluble COD of a properly designed and well operated biological treatment plant is composed of SP+SI1. Experimental protocols are developed for the identification of readily biodegradable COD, SS1; soluble inert COD, SI1; particulate inert COD, XI1; particulate inert microbial products, XP and soluble inert microbial products, SP under aerobic conditions (Ekama et al., 1986; Germirli et al., 1991 and 1993; Orhon et al., 1999a; Orhon and Okutman, 2003). Similar experimental procedures defining soluble inert COD, SI1; particulate inert COD, XI1; particulate inert microbial products, XP and soluble inert microbial products, SP under anaerobic conditions are also available in literature (Germirli Babuna et al., 1998a; Ince et al., 1998). Once readily biodegradable COD, SS1; soluble inert COD, SI1; and particulate inert COD, XI1 are experimentally assessed rapidly hydrolysable COD (soluble), SH1, and slowly hydrolysable COD (particulate), XS1 can be determined from mass balance equations.

3. INERT COD AS A PRACTICAL TOOL Valuable information of practical use can be derived by assessing the soluble residual COD components of wastewaters. These inert COD fractions either initially present in the wastewater itself and/or microbially generated through the course of the biological processes, dictate if discharge standards can be complied with a prescribed biological treatment. Furthermore, it is possible to address the appropriate type of biological treatment applicable to the wastewater under investigation. Inert COD fractions can also be used as a tool to evaluate the performance of an existing biological treatment system. For this purpose COD removal efficiencies achieved in the full scale treatment plants must be compared with inert COD data denoting the attainable levels of COD removals. A full scale treatment plant removal efficiency lower than that indicated by inert COD data can be attributed to improper design and/or operational problems. Furthermore assessing the soluble inert COD fractions of segregated industrial effluent streams can be quoted as a vital issue for industrial wastewater management. In many industrial operations a part of the various auxiliary chemicals added during manufacturing processes, may end up as constituents of various waste streams. Once the inert nature of segregated industrial effluent streams susceptible of carrying recalcitrants auxiliaries is confirmed by running inert COD experiments, a management strategy can be developed. According to the so called minimization at source philosophy, streams with highly recalcitrant character can be directed towards a specific treatment prior to be mixed with biodegradable ones. Or else, substitution of xenobiotic auxiliaries with high recalcitrance by ecochemicals having biodegradable nature can be considered as an alternative management strategy. It should be noted that previously, the attention was given to substitute the chemicals that generate streams with high pollutant loads by the ones that engender less concentrated discharges. However current understanding attributes special emphasis to recalcitrance and toxicity without adressing pollutant loads measured in terms of conventional parameters.

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The effluent inert COD levels obtained for different industrial wastewaters after passing through aerobic and anaerobic treatment are tabulated in Table 2. Inert COD contents of segregated industrial discharges carrying commonly used auxiliary chemicals are also given in Table 2. Table 2. The effluent inert COD levels for different industrial wastewaters under aerobic and anaerobic conditions Wastewater Source CT1 SI1/CT1 (type of treatment) (mg/l) (%) Wastewater originated from Solid waste transport station (aerobic) 29900 1 (anaerobic) 29900 1 Pulp and paper mill (aerobic) 21500 21 Pulp and paper mill (anaerobic) 13000 6 Antibiotic formulation (aerobic) 415 78 Antibiotic formulation (aerobic) 400 63 Brewery (anaerobic) 90000 Cheese Whey (aerobic) 60000 Citric Acid Plant (aerobic) 29300 6 Laying chicken industry* (aerobic) 6380 5 (anaerobic) 6380 6 Laying chicken industry** (aerobic) 9900 7 (anaerobic) 9900 14 Alcohol distillery*** (aerobic) 2350 6 (anaerobic) 2350 3 Corn wet mill (aerobic) 4680 (anaerobic) 4680 7 Segregated industrial stream containing Natural Tannin (aerobic) 1100 2 Synthetic Tannin (aerobic) 465 29 Ligninsulphonate (aerobic) 615 52 Naphtalanesulphonate (aerobic) 1180 87 *Chicks step **Laying step ***after passing through a chemical treatment

SI1+SP (mg/l)

Reference

1470 1370

Kutluay et al., 2007

5310

Germirli Babuna et al., 1998a

2240

Eremektar et al., 1998

330

Iskender et al., 2007

260

Tezgel et al., 2007

3600

Ince et.al., 1998

260

Germirli et al., 1993

3390

Germirli et al., 1993

350 700

Germirli Babuna et al., 1999b

980 1570

Germirli Babuna et al., 1999b

340 85

Eremektar et al., 1999b

410 630

Eremektar et al., 2002

100

Germirli Babuna et al., 2007a

190

Germirli Babuna et al., 2007a

340

Germirli Babuna et al., 2007c

1034

Germirli Babuna et al., 2007b

Biodegradation Characteristics of Wastewaters

273

The data presented in Table 2 indicates that an anaerobic treatment can be prescribed for the treatment of wastewaters from the alcohol distillery. On the contrary for the laying step of laying chicken industry, since the application of aerobic treatment yields a lower effluent residual COD level than a corresponding anaerobic one; an aerobic type of treatment must be adopted. Due to the generation of inert metabolic products, it is not possible to obtain an effluent COD level lower than 3600 mg/l when treating the brewery wastewater under investigation by means of an anaerobic system. An effluent quality lower than 340 mg/l of COD can not be obtained when an aerobic biological treatment is prescribed for the segregated industrial stream carrying the investigated lignosulphonate derivative. Quite a high percentage of the COD generated by naphtalanesulphonate carrying discharge, namely 87 %, must be regarded as biorecalcitrant in nature (Germirli Babuna et al., 2007b). Such a remarkably high recalcitrant fraction indicates the necessity of applying a partial chemical pre-treatment to this segregated industrial wastewater stream prior to letting it mixed with the other wastewater sources which will consequently pass through a conventional biological treatment. The data on natural and synthetic tannin formulations represent an example that accentuates the importance of biodegradability while substituting an auxiliary with another one (Germirli Babuna et al., 2007a). The COD contents (CT1) tabulated for natural and synthetic tannin display that application of synthetic tannin as an auxiliary chemical must be preferred since in this case less COD is introduced to the segregated effluent. On the contrary a correct evaluation can only be obtained from the data on inert soluble COD. When the segregated discharges containing both of the tannin formulations are comparatively appraised in terms of the lowest achievable COD levels after biological treatment, natural tannin is observed to have a lower residual COD of 100 mg/l.

4. COD FRACTIONS OF DIFFERENT WASTEWATERS Data on COD fractions can also be used as a solid source of information for modelling studies applied to biological treatment systems. Table 3 tabulates the results of experimental work carried out on COD fractionation of a wide range of industrial wastewaters and domestic sewage. The data given on COD fractionation can be used as a guide. Nevertheless it should be noted that occasionally the figures obtained even on the same industrial branch dealing with similar operations can show discrepancies. As can be seen from the table mills manufacturing cotton knit fabric generate effluents having SI1/CT1 ratios ranging from 7 % to 34 % probably due to the usage of various auxiliary chemicals with different biodegradation characteristics. Therefore a case wise appraisal handling the sui generis nature of industrial facilities is recommended to reach a sound result. Detailed information on how to adopt COD fractionation for modelling purposes can be found in literature (Orhon et al., 2009).

274

Fatos Germirli Babuna and Derin Orhon Table 3. COD fractions of various wastewaters

Wastewater Source

Parameter (mg/l) CT1 ST1 SS1

SI1

SH1

XS1

XI1

SI1/CT1 (%)

CS1/CT1 (%)

580

139

20

411

352

-

2

97

515

144

17

355

229

117

2

85

1320

340

147

833

ND

ND

10

455

220

32

203

160

72

5

85

465

25

55

417

230

50

7

87

1660

12

125

40

1485*

ND

7

1700

330

100

1270

700

-

4

96

1570

325

240

1005

340

-

13

87

1650

410

100

1140

290

-

5

95

1485

300

415

770

390

110

21

74

1900

420

170

1310

365

35

7

91

1165

330

260

575

288

17

18

81

675

110

320

245

ND

ND

34

1558

ND

317

ND

517

25

15

1210

187

289

734

708

62

15

82

890

118

247

525

227

63

21

74

1690

165

250

1275

598

112

10

85

1075

400

-

675

230

105

-

93

1298

434

214

649

724

263

9

79

1090

382

175

534

-

-

16

84

775

230

120

425

-

-

16

84

1140

380

30

730

1155

305

1

87

5490

145

4300

2400

Textile Organized Industrial District 932 Predominantly Textile (Orhon et al., 1999c) Organized Industrial District 860 Predominantly Textile (Ubay Cokgor et al., 1998) Wool/PES Knit Fabric 1445 (Orhon et al., 2000b) Wool/PES Knit Fabric 687 (Dulkadiroglu et al., 2002) Acrylic Fibre Carpet 775 (Yildiz et al., 2007) Polyamide Fibre Carpet 1890 (Yildiz et al., 2008) Cotton Denim Fabric ** 2400 (Germirli Babuna et al., 1998c) Cotton Denim Fabric ** 1910 (Orhon et al., 2001 a) Cotton Denim Fabric ** 1940 (Orhon et al., 2001 a) PES Knit Fabric 1985 (Germirli Babuna et al., 1998c) Cotton Knit Fabric 2300 (Germirli Babuna et al., 1999a) Cotton Knit Fabric 1470 (Germirli Babuna et al., 1998c) Cotton Knit Fabric 955 (Dogruel, 2000) Cotton Knit Fabric 2100 (Germirli Babuna et al., 1998b) Cotton Knit Fabric 1980 (Orhon et al., 2001b) Cotton Knit Fabric 1180 (Orhon et al., 2001b) Cotton/PES Knit Fabric 2400 (Germirli Babuna et al., 1998c) Integrated Dairy Processing 1410 (Orhon and Ubay Cokgor, 1997) Tannery Primary Treatment Effluent 2285 (Orhon et al., 1999b) Chemical Treatment Effluent 1090 (Orhon et al., 1999b) Chemical Treatment Effluent 775 (Orhon and Ubay Cokgor, 1997) Meat Processing Integrated Meat 2600 (Gorgun et al., 1995) Integrated Meat 7290 (del Pozo et al., 2003)

94

Biodegradation Characteristics of Wastewaters

275

Table 3. (continued) Wastewater Source

Parameter (mg/l) CT1 ST1 SS1 2490 1770 308

SI1 240

SH1 1220

XS1 685

XI1 35

2500

2045

45

410

500

-

1495

840

105

550

305

-

6775

5370

45

1360

1310

-

3120

735

77

2308

260

-

3192

1710

-

1482

418

190

370

40

56

174

197

63

Poultry (Eremektar et al., 1999a) Confectionery Plant I 3000 (Orhon et al., 1995) Plant II 1800 (Orhon et al., 1995) Plant III 8085 (Orhon et al., 1995) Plant IV 3380 (Orhon and Ubay Cokgor, 1997) Corn Wet Mill 3800 (Eremektar et al., 2002) Hotel 630 (Orhon and Ubay Cokgor, 1997) Domestic Sewage Raw 431 (Orhon et al., 1997) Settled 300 (Orhon et al., 1997)

ND: negligible settling

170

42

11

117

229

32

170

42

11

117

114

16

PES: Polyester

*soluble

SI1/CT1 (%) 10

CS1/CT1 (%) 89

2

98

6

94

1

99

2

98

-

95

9

81

3

90

4

91

** After passing through 6 hours of gravity

5. CONCLUSION COD fractionation providing a new dimension for the biodegradation characteristics of different fractions is the most useful experimental approach today for the elucidation of wastewater character and the engineering decisions for appropriate treatment. This approach also serves as an integral part of currently used mechanistic models. Recognition and expermental assessment of the recalcitrant portions of COD in wastewaters, together with the discovery of residual metobic products generation have to be considered as an important milestone for performance evaluation of applicable biological treatment processes. However, it is essential to recognize that COD fractionation may become more detrimental than beneficial, if not properly assessed with the necessary understanding of relevant microbial mechanisms associated with different experimental techniques and full account of various analytical uncertainties.

ACKNOWLEDGMENT Reviewed by Prof. Orhan YENIGUN Affiliation of the reviewer: Director, Bogazici University, Institute of Environmental Sciences

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REFERENCES Andreottola, G., Foladori, P., Ragazzi, M. and Villa, R. (2002) Dairy wastewater treatment in a moving bed biofilm reactor. Water Sci. Technol. 45(12), 321-328. Arslan Alaton, I., Insel, G., Eremektar, G., Germirli Babuna, F. and Orhon, D. (2006a) Effect of textile auxiliaries on the biodegradation of dyehouse effluent in activated sludge. Chemosphere. 62(9), 1549-1557. Arslan Alaton, I., Okay S.O., Eremektar, G. and Germirli Babuna, F. (2006b) Toxicity assessment of raw and ozonated textile biocides. International Magazine for Textile Design Processing and Testing, AATCC Rev. 6(5), 43-48. Arslan Alaton, I., Iskender, G., Ozerkan, B., Germirli Babuna, F. and Okay, O. (2007) Effect of chemical treatment on the acute toxicity of two commercial textile dye carriers. Water Sci. Technol. 55(10), 253-260. Chamarro, E., Marco, A. and Esplugas, S. (2001) Use of fenton reagent to improve organic chemical biodegradability, Water Res. 35(4), 1047-1051. del Pozo, R., Tas, D.O., Dulkadiroglu, H., Orhon, D. and Diez, V. (2003). Biodegradability of slaughterhouse wastewater with high blood content under anaerobic and aerobic conditions. J. Chem. Technol. Biot. 78(4), 384-391 Dogruel, S. (2000) The Effect of Ozonation on COD Fractions - A Case Study for Cotton Finishing ill. M.Sc. Thesis, Institute of Science and Technology, Istanbul Technical University, Istanbul (in Turkish). Dulkadiroglu, H., Eremektar, G., Dogruel, S., Uner, H., Germirli Babuna, F. and Orhon, D. (2002) In-plant control applications and their effect on treatability of a textile mill wastewater. Water Sci. Technol. 45(12), 287-295. Ekama, G.A., Dold, P.L., and Marais, G.v.R. (1986) Procedures for determining influent COD fractions and the maximum specific growth rate of heterotrophs in activated sludge systems. Water Sci. Technol. 18, 91-114. Eremektar, G., Germirli Babuna, F. and Orhon, D. (1998) Inert COD fractions in two-stage treatment of a pulp and paper mill effluent, J. Chem. Technol. Biot. 72, 7-14. Eremektar, G., Ubay Cokgor, E., Ovez, S., Germirli Babuna, F. and Orhon D. (1999a) Biological treatability of poultry processing plant effluent-a case study. Water Sci. Technol. 40, 323-329. Eremektar, G., Germirli Babuna, F. and Ince, O. (1999b) Fate of inert COD fractions in twostage treatment of a strong wastewater. J. Environ. Sci. Heal. A. 34(6), 1329-1340. Eremektar, G., Karahan Gul, O., Germirli Babuna, F., Ovez, S., Uner, H. and Orhon, D. (2002) Biological treatability of a corn wet mill effluent, Water Sci. Technol. 45(12), 339-346. Grady, C.P.L., Daigger, G.T. and Lim, H.C. (1999) Biological Wastewater Treatment: 2nd edition, Marcal Dekker Inc., New York. Green, M., Sheler, G. and Moraine, R. (1981) Chemical and biochemical oxygen demand as indicators of biodegradable substrate concentration. J. Water Pollut. Control Fed. 655658. Germirli, F., Orhon, D. and Artan, N. (1991) Assessment of the initial inert soluble COD in industrial wastewaters. Water Sci. Technol. 23(4-6), 1077-1086.

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Germirli, F., Orhon, D., Artan, N., Ubay, E. and Gorgun, E. (1993) Effect of two-stage treatment on the biological treatability of strong industrial wastes. Water Sci. Technol. 28(2), 145-154. Germirli Babuna, F., Ince, O., Orhon, D. and Simsek, A. (1998a) Assessment of inert COD in pulp and paper mill wastewater under anaerobic conditions. Water Res. 32(11), 34903494. Germirli Babuna, F., Eremektar, G., and Yaprakli, B. (1998b) Inert COD Fractions of Various Textile Dyeing Wastewaters. Fresen. Environ. Bull. 7, 959-966. Germirli Babuna, F., Orhon, D., Ubay Cokgor, E., Insel, G. and Yaprakli, B. (1998c) Modelling of activated sludge for textile wastewaters. Water Sci. Technol. 38(4-5), 9-17. Germirli Babuna, F., Soyhan, B., Eremektar, G. and Orhon, D. (1999a) Evaluation of treatability for two textile mill effluents. Water Sci. Technol. 40(1), 145-152. Germirli Babuna, F., Cekyay, E., Eremektar, G. and Orhon, D. (1999b) Pollution loads and inert COD in laying chickens industry. Water Sci. Technol. 40, 207-213. Germirli Babuna, F., Yilmaz, Z., Okay, O., Arslan Alaton, I and Iskender, G. (2007a) Ozonation of synthetic versus natural tannin: inert COD and toxicity towards Phaeodactylum tricornutum. Water Sci. Technol. 55(10), 45-52. Germirli Babuna, F., Camur, S., Arslan Alaton, I., Okay, O. and Iskender, G. (2007b) The application of ozonation for the detoxification and biodegradability improvement of a textile auxiliary: naphthalene sulphonic acid. Proceedings of the 10th International Conference on Environmental Science and Technology, Kos Island, Greece, A-402-409. Germirli Babuna, F., Oructut, N., Arslan Alaton, I., Iskender, G. and Okay, O. (2007c) Reducing the toxicity and recalcitrance of a textile xenobiotic through ozonation. International Conference on Environmental Survival and Sustainability. 19-24 February 2007, Near East University, Nicosia, Northern Cyprus. (provisionally accepted for publication in J. Environ. Sci. Heal. A) Gorgun, E., Ubay Cokgor, E., Orhon, D., Germirli, F. and Artan, N. (1995) Modelling biological treatability for meat processing effluent. Water Sci. Technol. 32(12), 43-52. Ince, O., Germirli Babuna, F., Kasapgil, B. and Anderson, G.K. (1998) Experimental determination of the inert soluble COD fraction of a brewery wastewater under anaerobic conditions. Environ. Technol. 19, 437-442. Iskender, G., Sezer, A., Arslan Alaton, I., Germirli Babuna, F. and Okay, O.S. (2007) Treatability of Cefazolin antibiotic formulation effluent with O3 and O3/H2O2 processes. Water Sci. Technol. 55(10), 217-225. Koyunluoglu, S., Arslan Alaton, I., Eremektar, G. and Germirli Babuna, F. (2006) Preozonation of commercial textile tannins: effect on biodegradability and toxicity. J. Environ. Sci. Heal. A. 41(9), 1873-1886. Kutluay, G., Iskender, G., Germirli Babuna, F. and Orhon, D. (2007) Treatment options for effluents from a solid waste transport station, Desalination 211, 96-101. Orhon, D., Artan, N., Buyukmurat, M. and Gorgun, E. (1992) The effect of residual COD on the biological treatability of textile wastewater. Water Sci. Technol. 26(3-4) 815-825. Orhon, D., Gorgun, E., Germirli, F. and Artan, N. (1993) Biological treatability of dairy wastewaters. Water Res. 27(4), 625-633. Orhon, D., Yildiz, G., Ubay Cokgor, E. and Sozen, S. (1995). Respirometric evaluation of the biodegradability of confectionary wastewaters. Water Sci. Technol. 32(12), 11-19.

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Orhon, D. and Ubay Cokgor, E. (1997). COD Fractionation in wastewater characterization – the state of the art. J. Chem. Technol. Biot. 68, 283-293. Orhon, D., Ates Genceli, E., Sozen, S. and Ubay Cokgor, E. (1997) Characterization and COD fractionation of domestic wastewaters. Environ. Pollut. 92(2), 191-204. Orhon, D., Karahan, O. and Sozen, S. (1999a). The Effect of microbial products on the experimental assessment of the particulate inert COD in wastewaters. Water Res. 33(14), 3191-3203. Orhon, D., Ates Genceli, E. and Ubay Cokgor, E. (1999b) Characterization and modeling of activated sludge for tannery wastewater. Water Environ. Res. 71(1), 50–63. Orhon, D., Taşlı, R. and Sözen, S. (1999c). Experimental Basis of Activated Sludge Treatment for Industrial Wastewaters - The State of the Art, Water Sci Technol, 40, 1, 1 11 Orhon, D., Sozen, S., Kabdasli, I., Germirli Babuna, F., Karahan, O., Insel, G., Dulkadiroglu, H., Dogruel, S., Kiran, N., Baban, A. and Kemerdere Kaya, N. (2000b) Recovery and reuse in the textile industry – A case study at a wool and blends finishing mill. In: Chemical Water and Wastewater Treatment VI, ed. H. H. Hahn, E. Hoffman and H. Odegaard, 305-315, Springer Verlag, Berlin. Orhon, D., Germirli Babuna, F. and Insel, G. (2001a) Characterization and modelling of denim processing wastewaters for activated sludge. J. Chem. Technol. Biot. 76(1), 1-13. Orhon, D., Germirli Babuna, F., Kabdasli, I., Insel, G., Karahan, O., Dulkadiroglu, H., Dogruel, S., Sevimli, F. and Yediler, A. (2001b) A scientific approach to wastewater recovery and reuse in the textile industry. Water Sci. Technol. 43(11), 223-231. Orhon, D. and Okutman, D. (2003). Respirometric assessment of residual organic matter for domestic sewage. Enzyme Microb. Tech. 32(5), 560-566. Orhon, D., Germirli Babuna, F., and Karahan, O. (2009). Industrial Wastewater Treatment by Activated Sludge. IWA Publishing, ISBN: 9781843391449. Sharma, K.P., Sharma, S., Subhasini Sharma Singh, P.K., Kumar, S., Grover, R. and Sharma, P.K. (2007) A comparative study on characterization of textile wastewaters (untreated and treated) toxicity by chemical and biological tests. Chemosphere. 69, 48-54. Tezgel, T., Germirli Babuna, F., Arslan-Alaton, I., Iskender, G. and Okay, O. (2007) Pretreatment of Ceftriaxone formulation effluents: drawbacks and benefits. International Conference on Environmental Survival and Sustainability. 19-24 February 2007, Near East University, Nicosia, Northern Cyprus. (provisionally accepted for publication in J. Environ. Sci. Heal. A) Ubay Cokgor, E., Sozen, S., Orhon, D. and Henze, M. (1998) Respirometric analysis of activated sludge behaviour-1. assessment of the readily biodegradable substrate. Water Res. 32(2), 461-475. Yildiz, G., Insel, G., Cokgor, E.U. and Orhon, D. (2007) Respirometric assessment of biodegradation for acrylic fiber based carpet finishing wastewaters. Water Sci. Technol. 55(10), 99-106. Yildiz, G., Insel, G., Cokgor, E. U. and Orhon, D. (2008) Biodegradation kinetics of slowly biodegradable substrate in polyamide carpet finishing wastewater, J. Chem. Technol. Biot. 83, 34-40.

In: Fluid Waste Disposal Editor: Kay W. Canton, pp. 279-303

ISBN: 978-1-60741-915-0 © 2010 Nova Science Publishers, Inc.

Chapter 13

BATCH TREATMENT OF A COFFEE FACTORY EFFLUENT FOR COLOUR REMOVAL USING A COMBINATION OF ELECTRO-COAGULATION AND DIFFERENT SUPPORTING ELECTROLYTES L. Etiégni1*, D. O. Oricho1, K. Senelwa1, B. O. Orori1, B. K. Balozi1 K. Ofosu-Asiedu2 and A. Yitambé3 1

Department of Forestry and Wood Science, Moi University, P. O. Box 1125 Eldoret, Kenya. 2 J.I.C., Dept. of Chem. Eng. Box 10099, Jubail Industrial City-31961, Kingdom of Saudi Arabia. 3 Kenyatta University, Department of Public Health P.O Box 43844-00100 Nairobi, Kenya

ABSTRACT In the present study, two types of colour removal systems were tested on effluent samples collected from a coffee pulping factory which discharged on average 15 m3 of wastewater daily with a colour index of about 2500 OH that was too high for direct discharge into a river in Kenya. The two colour removal systems used were: (i) electrolysis combined with wood ash or coffee husks leachate and (ii) electrolysis combined with phosphate rock solutions at a rate of 0.5 g/l to 4g/l. Phosphate rock is often used as agricultural liming agent. The surface area of the electrodes was set at close to 75 m2/m3 of effluent with a current density of 1,200 mA/m2. The experiments were laid out in a stratified random sampling design and the data were analysed using the Statistical Package for Social Scientists (SPSS) computer programme version 10.0. Electrolysis combined with phosphate rock (ELPHOS) proved to be the best process in terms of power consumption (68% reduction) compared with the 57% reduction by electrolysis combined with wood ash (ELCAS) and the 58% reduction by electrolysis combined with *

Corresponding author: Email: [email protected]

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L. Etiégni, D. O. Oricho, K. Senelwa et al. coffee husks ash (ELCHAS). Besides the 100% colour removal, ELPHOS also reduced other effluent physico-chemical parameters such as BOD, COD, TSS and TS by 79%, 80%, 69%, and 88% respectively. The analysis of ELPHOS treated wastewater showed that the mill could discharge an effluent that meets local discharge standards for colour requirements. It is recommended that recycling of the treated water by ELPHOS back to the factory for cleaning and washing purposes be considered since the quality meets the requirement for uses of fresh water for cleaning purposes. Furthermore, calculation of power consumption based on a scale-up batch reactor of 15 m3 proved less expensive to treat the factory effluent than a set of 12 one 100-L reactors similar to the one used in the field.

Keywords: Coffee; colour; wastewater: pulping; phosphate rock; electrochemical method.

INTRODUCTION Coffee is a member of the large Rubiacea family, where it constitutes the coffea genus (Coste, 1992). The genus coffea contains approximately 70 species, and the most widely grown are the arabica and the robusta species (Cambrony, 1992). Today, after petroleum, coffee euphemistically referred to as ―Black gold‖, is the World's most important traded commodity, standing above coal, meat, wheat and sugar. The global harvest of coffee, however, is subject to considerable fluctuations from year to year. These fluctuations are caused by a variety of factors such as climate-induced fluctuations especially in Brazil, the World larger coffee producer, the amount of coffee produced and the price charged which are determined by the commercial policy interests of the producing and purchasing countries. Its cultivation, processing, trading, transportation and marketing provide employment for a large population base in all producing countries (Muleta, 2007). Botanical evidence indicates that coffee plant "Coffea arabica", the variety mostly grown in Kenya originated from the Abyssinian highlands of central Ethiopia where it was cultivated by Harrrar tribe but can still grow wild today. Somehow Arab traders got the beans from Ethiopia across the Red Sea to Yemen around the 6th century AD. Before this, African Indigenous populations in Ethiopia were using the beans as a solid food: the ripe berries were squashed, combined with animal fats and shaped into round balls that could be carried and eaten on long journeys. It wasn't until 1615 that the first shipment of coffee arrived in Europe at Venice (the European trading headquarters at that time) from Turkey, and coffee houses quickly spread through Italy to Vienna (in today Austria), then on to most of Europe. France is credited with the first introduction of coffee to the Americas through its colonies of the Martinique, the West Indies and the French Guiana where the first coffee plantations were founded in 1720. In 1727, from these French colonies, coffee found its way to Brazil, where it became the mainstay of the economy, accounting for 63 percent of the country's exports by 1891. In 1893 coffee from Brazil was introduced to the then British colonies of Kenya and Tanganyika in East-Africa, only a few hundred kilometres south of where it had originated, in Ethiopia. Today coffee is grown in most parts of Kenya employing around 6 million people across a country of about 40 millions inhabitants where it constitutes one of the two major cash crops as well as an important export commodity, accounting for 15% of total export and contributing substantially to the country‘s economy. In 2005 for example, Kenya exported coffee worth a total of US$ 131 million. Kenya coffee beans are considered by gourmet

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coffee lovers as some of the finest in the world. They are ranked with beans from Jamaica's Blue Mountain or Hawaii's Kona region for quality and flavour. Coffee making can be divided into two types of processes: green coffee and dried coffee beans processing. Water is mainly used during washing of the green coffee especially if the process is a wet one. Wet-processing coffee is a relatively new method of removing the four layers surrounding the coffee bean. This process results in a coffee that is cleaner, brighter, and fruitier. Most countries (e.g. Kenya) with coffee valued for its perceived acidity, process their coffee using the wet-process (Coffee Research Institute, 2001). In SOCIFINAF, a large coffee estate in Kenya with a coffee processing factory at Mchana, near Nairobi where the experiment on colour removal was conducted, wet processing is used, which involves the following steps: the removal of pulp (pulping) and mucilage (mucilage removal), washing and fermentation for fully washed coffee as shown in Figure 1. During pick production, the coffee estate at Mchana factory collects roughly 60,000 kg of green beans per day for a final output of 3000 kg per day of dry coffee beans ready for export. On average, 15 m3 of wastewater are generated per day from this factory. Transportation to Factory

Drying

Pulping

Washing

Washing

Removal of mucilage

Figure 1. Production stages in a Coffee Factory using wet processing. Washing of the fermented coffee which is made of pectin materials has a negative impact on the effluent parameters. The pectin materials include protopectin (33%), reducing sugars namely glucose and fructose (30%), non-reducing sugars such as sucrose (20%), cellulose and ash (17%). After thorough straining, the mucilage is allowed to soften and decompose before its removal through washing which is generally carried out in large tanks, called washers, by slowly moving coffee through open-topped channels (Coste, 1992). Wastewater from fermentation and washing stages is kept in a concrete embankment where solids are allowed to settle over night before the effluent is released into a near-by river. At Mchana factory, the treated effluent still has a high colour level of 2,500 oH and a high BOD of 266 mg/L that must be substantially reduced before any discharge into a near-by river is done according to new environmental regulations in Kenya. Colour can damage aquatic life in a river through the inhibition of light penetration into water and may adversely affect the aesthetical value of any body of water. In addition, unlike the removal of BOD, COD, suspended solids, and toxicity, which can be successfully reduced through primary treatment followed by biological treatment, colour is extremely difficult and expensive to remove (UNEP, 1981; Ikehata and Buchanan, 2000). Therefore there was a need to find an economical method of reducing these effluent parameters to conform to local discharge standards. There are several methods of removing colour which include: adsorption, soil media, coagulation, ultraviolet irradiation, and membrane-based technology such as ultrafiltration

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(ASTM,1983; Prasad and Joyce, 1991; Ikeheta and Buchanan, 2000). These technologies can reduce pollution load and effluents colour to acceptable levels, but most are quite expensive and very few are in common practice. Electrochemical treatment alone or combined with wood ash leachate (ELCAS) has recently been tested and proved to be an effective method of colour reduction from a pulp and paper mill (Etiégni et al. 2004; Orori et al, 2005). However, over electrical polarisation within electrodes, which normally causes excess voltage with wastage of electrical energy remains the biggest impediment of this method (Springer et al., 1995; Ahonen, 2001). Electrical power cost is as high as US¢25.0 per kWh in Kenya, more than six times what it is in South Africa, the continent‘s industrial power house. In addition, high pH, high sludge production, high electrical conductivity of the treated effluent and unavailability of wood ash when needed at some factories may hamper the successful application of electrochemical methods. Good alternatives to wood ash that will not only reduce power consumption during electrochemical application, but also be inexpensive must be identified. In this paper, we report on the potential use of phosphate rock and coffee husks ash as substitute for wood ash during the removal of colour from a green coffee processing mill effluent by electrochemical method.

MATERIALS AND METHODS The Study was conducted on a coffee factory effluent. The company is one of the largest in Kenya and produces from this factory on average about 15m3 of effluent daily from water obtained from wells or dam found within the factory premises. DC power supply

Non-conducting material Iron electrodes +ve

-ve

47cm 47cm

30cm 45cm

65cm

Figure 2. Batch reactor for colour removal from a coffee factory effluent.

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Whenever possible, at least three replications were obtained. Chemical analysis of the raw and treated coffee factory effluent was carried out for quality assessment using AAS or flame photometer. Data were analyzed using Analysis of variance, Least significant difference and Duncan test between different treatments using SPSS statistical package. Samples of wastewater were collected at the effluent discharge point and promptly transported for preservation at low temperatures (below 4oC) to avoid biological degradation before analysis (UNESCO/WHO, 1978; Arudel, 2000). Temperature was measured at the point of collection. Several other parameters such as alkalinity, colour, turbidity, pH, biochemical oxygen demand, chemical oxygen demand, total solids, total suspended solids and electrical conductivity were determined according to TAPPI standard procedures (Tappi Test Methods, 1992). Colour of the effluent was measured by use of a Loviband colour comparator. To Power Supply Non conducting material

48.0 cm Iron Electrodes 2.0 mm thick

30.0 cm Figure 3. Electrodes configuration.

The electro-coagulation experiments were carried out with wood ash (ELCAS), phosphate rock (ELPHOS) and coffee husks ash (ELCHAS) in a set-up depicted in Figure 2. Experiments were performed in a monopolar batch reactor, with eighteen sacrificial iron electrodes connected in parallel (monopolar parallel mode) and kept 20.0 mm apart using a non-conducting material (Figure 3). Only the outer electrodes were connected to the power source and anodic and cathodic reactions occurred when the current passed through the electrodes. The size of the internal plastic cell was 45 x 65 x 47 cm (width × length × depth) with an effective volume of 100 litres. The electrodes were immersed 7/8 deep into the effluent sample to achieve surface area coverage of 75 m2/m3 of wastewater. A power supply pack with an input of 220 V and variable output of 0-40 V with maximum current of 1.5 amperes was used as a direct current source during electrolysis. On average, the current density was maintained at around 1200 mA/m2. After each run, the iron electrodes were rinsed in a water

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bath following a scrubbing with a piece of cloth to minimize electrode fouling. For supporting electrolytes, leaching of various materials i.e. wood ash, coffee husks ash and phosphate rock was first carried out for 3hrs or 12 hrs (overnight soaked). After that, different volumes of the leachates were used to yield concentrations of 500, 1000, 1500, 2000, 2500, 3000, 3500, and 4000 g/m3 of wastewater. For each run, the power required for complete colour removal (0oH) was measured and power consumption calculated using Equation 1. Phosphate rock (MPR) used in this work came from Northern Tanzania and was chosen because of its reactivity and availability on local market as a low cost fertilizer. MPR chemical concentration has been reported by van Kauwenbergh (1991). Wood ash and coffee husks ash were obtained from the coffee factory. Power (Watts.hr) = Current (I) * Potential Difference (V) * Time (hr) RESULTS AND DISCUSION

(1).

At the start of each run, the coffee factory effluent was slightly brown. However, with the application of the current, the effluent turned darker and became more opaque as the reaction progressed. A layer of foam started forming at the surface of the wastewater probably due to the production of hydrogen gas. This was followed by the formation of a cloud of flocs and subsequent rapid decantation.

10000 R/water COD for NSC

COD for OSA

COD for NSA

COD for OSR

COD for NSR

COD for OSC

1000

COD (mg/l)

100

10

1 0

500

1000

1500

2000

2500

3000

3500

4000

Concentration (g/m3) Figure 4. COD reduction for different treatments.

For this study, the formation of a cloud was selected as an indication of the end of the electrocoagulation reaction. Effluent temperature was normal at around 20oC. Before colour removal, the effluent had a BOD of 851 mg/l, COD of 1845 mg/l and a colour of 2500 oH. These effluent characteristics were found to be well above the 30 mg/l, 50 mg/l and 15 oH for BOD, COD and colour respectively, the expected limits set by the National Environment Management Authority (NEMA); the environmental regulatory body in Kenya.

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The pH of 5.2 was also lower than the recommended 6.5. The results of treated effluent parameters for the coffee factory are listed in Table 1. ELCAS, ELCHAS and, ELPHOS had the same reduction range for TS and TSS in the treated effluent from 85 to 89%. This reduction was, however, not significantly affected by any increase in the concentration of wood ash, coffee husks ash or phosphate rock leachate. The reduction of solids was probably due to the fact that the two compounds contained comparable amounts of Ca2+ that acted as a counter-ion during the experiment. Figure 5 shows BOD reduction of selected treatments and soaking times.

Raw water ELCAL ELCAS SH3 ELCAL ELPHOSSH3

ELCAS BH6 ELPHOS BH6

BOD (mg/l )

300 200 100 0

Treatment Figure 5. BOD of selected treatments and of soaking time

The COD values for treated effluent reduced by ELCAS, ELCHAS and ELPHOS (for both overnight soaked (OS) and not overnight soaked (NS)) are shown in Figure 4. Compared to the initial raw effluent, the factory‘s COD decreased by between 66 and 91% , 51 and 83 % and 68 and 94 % through ELCHAS, ELCAS and ELPHOS methods respectively. ELPHOS was more effective for reducing COD than ELCAS, and the difference was statistically significant at P = 0.05. The final COD of 114-594 mg/l was still higher than the local effluent discharge standard requirements. With the increase in the SE leachate concentration, COD and BOD values decreased for ELCAS SH3, ELPHOS BH6, and ELCAS BH6. The reduction of COD during this experiment was slightly higher than the removal rate of 74.0% reported by Orori (2003) and Orori et al. (2005). The effect of soaking time on COD reduction through ELCHAS, ELCAS or ELPHOS is shown in Figure 4. The results indicate that as wood ash, coffee husks ash and phosphate rock were allowed to soak overnight, more removal of COD occurred in the subsequent experiment. The explaination is that with increase in soaking time, more supporting electrolyte (SE) was made available in the reactor, thereby increasing the electrical conductivity of the solution in most cases and helping the electrocoagulation process (WeiLung et al, 2009). Unlike with wood and coffee husks ashes, leaching from phosphate rock reduced the electrical conductivity of the factory‘s effluent (Table 1). The increased efficiency of phosphate rock leachate cannot therefore be explained simply through increased electrical conductivity.

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30 OSA

NSA

OSC

NSC

OSR

NSR

25

Power (Whr)

20

15

10

5

0 0

1000

2000

3000

4000

5000

Electrolyte Dosage (g/m 3 )

Figure 6. Effect of supporting electrolytes on electrocoagulation power consumption.

Soaking time seemed to help the reduction of BOD although for example, the difference between soaking time of 3 and 6 hours was not statistically significant (P> 0.05). This confirmed findings by Etiégni and Campbell (1991) that generally during wood ash leaching experiments, more than 90% of materials is expected to leach out after 30 minutes. ELCHAS reduced BOD by between 47 and 88%, ELCAS between 28 and 78%, while ELPHOS removed BOD between 51 and 90% and the differences were statistically significant (P< 0.05), although the wide variation in the recorded BOD removal values could not be easily explained. Wood ash was the primary source of hydroxide ions that enhanced the process for colour and BOD removals and probably led to reduced power consumption. Low power consumption could also be attributed to the catalytic properties of metal oxides such as MgO Mn2O3, Cr2O3, PbO2 found in wood ash and phosphate rock. The catalytic properties of these metal oxides on the surface, or in the space between the anode and cathode during electrocoagulation has been recognised in previous studies and might have assisted the reduction of the time for current flow as reported by Ahonen (2001). Reduction of BOD by ELPHOS in this experiment was in some instances higher than the values reported by both Springer et al. (1995) of 70% and Ahonen (2001), and yielded a final BOD value of between 20 and 40 mg/l, the higher end of which is not acceptable by local discharge standards.

Power Reduction Electrolysis combined with phosphate rock (ELPHOS) proved to be the best in terms of power consumption (68% reduction) compared to ELCAS (57% reduction) and ELCHAS (58% reduction), and the differences were statistically significant (P< 0.05) although it also depended on the quantity of supporting electrolyte (SE) added to the effluent. The effect of various SE concentrations on the treatment efficiency and power reduction of coffee factory

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effluent is shown in Figure 6. Wood ash, coffee husks ash and phosphate rock used as supporting electrolytes helped reduce power consumption (Figure 6). Soaking seemed to avail more supporting electrolyte to the colour removal process, although this may not be true with phosphate rock. Of the three supporting electrolytes, coffee husks ash had the least impact on colour removal when not soaked, followed by wood ash and phosphate rock. The latter showed no significant difference between soaked and not soaked supporting electrolyte. This stems from its low solubility which has been observed in studies where phosphate rock was used in agriculture for pH control. When the concentration of the supporting electrolyte (SE) in solution increased, power consumption reduced probably as a result of increased conductivity (Figure 7). If the required voltage of the electrocoagulation reaction is expressed as follows: where EC EA δA δC IRcell IRcircuit

V = EC - EA- δA- δC - IRcell - IRcircuit

(2)

= Electrical Potential at the cathode = Electrical potential at the anode = Zeta potential at the anode = Zeta potential at the cathode = voltage of the electrocoagulation cell = voltage of the electrical circuit

The removal of colour may have resulted from the combined effect of Fe2+ and Fe3+ generated in the solution during electrolysis of iron electrodes (Othman et al., 2006). It took on average between 33 to 54 minutes, depending on the SE concentration, for the effluent colour to be completely removed. These results could help determine the reactor‘s volume needed based on the detention time for complete decolourization of the coffee factory‘s effluent, if there is need for a continuous flow reactor:

Volume Detention time = Effluent flow rate 0

165

330

500

665

830

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Power (kW/Yr)

4.0 3.5 3.0 2.5 2.0 1.5 1.0 0.5 0.0

ELCAS BH6 ELCAS SH3 ELPHOS BH6 ELPHOS DH3

ELCAL

Figure 7. Power consumption for different treatment system and SE concentrations.

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It shows that the necessary voltage v to access a certain current density had reduced because of the introduction of the SE, so that the consumed electrical energy had also decreased (Kashefialasl et al., 2006). SE is said to compress the double layer which in turn reduces the Zeta potential of the substrate ions and helps their agglomeration or coagulation. The reduction of power consumption in this experiment was slightly lower than the 80% reduction reported by Etiegni et al. (2005) or Orori et al. (2005), probably because of the higher electrode surface coverage of 80 m2/m3 used in these experiments. Higher surface coverage normally lead to high efficient color removal. The presence of metal ions such as Ca, Fe, Al may have also helped the coagulation process. Chemical analysis of wood ash in previous studies has showed a significant presence of several chemical elements such as Ca and Fe in the form of their corresponding oxides which can act as coagulant when dissolved in water (Etiégni and Campbell, 1991).

Color Removal Colour was effectively removed from the coffee factory‘s effluent through electrocoagulation method. There appeared to be a positive effect of supporting electrolyte as ELCHAS, ELCAS and ELPHOS removed 100% colour confirming the effectiveness of this SE in electrocoagulation (Prasad and Joyce, 1991; Koparal and Gütveren, 2002). ELCHAS, ELCAS and ELPHOS had a negative effect on the treated effluent pH as it increased by between 27 and 75% for ELCHAS, 19 and 47% for ELCAS and between 9 and 22% for the ELPHOS as shown Table 1. The pH from ELPHOS was lower than ELCAS probably because of the slow reactivity of phosphate rock compared to wood ash. ELCHAS had a much higher impact on the coffee effluent final pH. Aeration had the effect of reducing the factory treated effluent COD (Figure 8). Using ELCAS followed by over-night aeration, COD was reduced to almost 5 mg/l after an initial spike to 68 mg/l, while with ELPHOS, COD of the treated effluent remained almost constant at 45 mg/l. This shows that after treatment with ELCAS, normal aeration often carried out in most wastewater treatment processes could further reduce the coffee factory wastewater parameters.

ELPHOS ELCAS

COD (mg/l)

80 60 40 20 0 1

2

3

4

5

Time (hr) Figure 8. COD reduction as a function aeration time.

6

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Total Solids There was a considerable reduction in solids in the treated effluent from the coffee factory. The reason for such behaviour could be due to the fact that the EC treatment may have induced the settling velocity of the suspended particles in which more suspended particle agglomerates cloaked together. The exposure of wastewater to EC treatment would contribute to a greater ionic charge so that more particles would collide and this would eventually help in enhancing particles‘ attraction and agglomeration (Othman et al., 2006). Table 1. Wastewater parameter for raw and treated effluent followed by aeration (with ash socked for six hours). Parameter Temperature (oC) pH TS (mg/l) TDS (mg/l) TSS (mg/l) TVS (mg/l) COD (mg/l) BOD5 (mg/l) EC (μΩ/cm) Colour (0H) Alkalinity (mg/l)

Effluent discharge Standards Kenya < 30 6.5-8.5 1200 30 50 30 15 -

Raw effluent 19 5.2 1920 400 1520 1070 1845 851 556 2500 129

ELCHAS

ELCAS

ELPHOS

20

20

20

6.6-9.1 220-300 40-160 120-220 160-300 164-624 102-456 705-773 0 149-202

5.3-7.9 220-340 40-180 140-180 160-320 320-899 186-616 508-677 0 114-156

5.9-6.8 220-300 40-160 120-180 80-180 114-594 86-420 197-270 0 55-76

Table 2. Chemical Analysis of raw and treated coffee factory’s effluent. Chemical element Cd Co Cr Cu Fe K Mg Mn Na P Pb Zn

Raw effluent 0.058 0 0.506 0.016 1.014 0.089 8.132 0.150 4.40 3.56 0.08 0.0273

Concentration (mg/l) ELCAS ELPHOS effluent effluent 0.005 0.002 0 0 0.100 0.506 0 0.026 0.203 0.006 9.0 0.017 0.265 2.094 0.4518 0 440.0 0.59 0 2.72 0 0.07 0.2352 0.0708

Effluent Standards 0.01 2.0 1.0 0.01 0.5

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Best concentration (mg/L) 4000 4000 3000

Amount to be used per year (tonnes) 24 24 18

R/water

OSA

Power consumption per m3 (WH) 42 48 42

NSA

OSC

Power consumption per year (KWH) 252 288 252

NSC

OSR

Total cost for treatment $/ year 6816 2504 2216

NSR

1400 1200 1000 EC

800 600 400 200 0 0

1000

2000

Concentration

3000

4000

(g/m 3)

Figure 9. Effect of SE on electrical conductivity of coffee factory effluent.

Effluent electrical conductivity (EC) and alkalinity results are also described in Table 1. The results show that ELCHAS increased EC by between 27 and 39%, ELCAS process increased slightly the effluent EC by between 8.7 and 21%, while ELPHOS reduced EC by as much as 51% and their net effect were statistically significant. An analysis of treated effluent in Table 2 shows that ELPHOS did not substantially increase minerals, even P for which it is used in agriculture. However ELCAS increased Na and Mn concentrations in the treated effluent making it unfit for certain uses such as agricultural irrigation.

R/water

OSA

NSA

OSC

NSC

OSR

NSR

Alkalinity (eq/m 3)

600 500

400 300 200

100 0 0

1000

2000

3000

Concentration (g/m 3)

Figure 10. Effect of SE on the effluent alkalinity.

4000

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Electrical Conductivity As mentioned earlier, the electrical conductivity of the treated effluent increased with the volume of supporting electrolyte. There seemed however to be a maximum in this increase which shows that maximum EC was achieved at 2000 mg/l or g/m3 except for never-soaked ash which exhibited a maximum EC at 4000 mg/l (Figure 9). The effect of phosphate rock on EC was negative. The difference between 2000 and 4000 mg/l was statistically significant (P< 0.05). Because of the negative impact of phosphate rock on EC, its overall positive effect on the removal of colour and other coffee effluent parameters cannot be explained through its impact on EC. Other factors such as the presence of CaO and MgO in phosphate rock may have played a role in helping treat the factory‘s effluent (van Kauwenbergh, 1991).

Alkalinity Alkalinity from ELCHAS increased from 15.0 to 56.0%, while ELPHOS first reduced alkalinity by as much as 59% (Table 1, Figure 10). The difference in the effect of these two processes was statistically significant (P< 0.05). The effect of ELCAS on coffee factory‘s effluent alkalinity was almost insignificant. The alkalinity value of the treated coffee factory effluent indicates that on average, the treated waste water could be more amenable to biological treatment if there was further need to improve the effluent characteristics prior to its discharge.

COST OF TREATMENT Assuming a discharge of effluent per year of 6000 m3, the average power consumption for selected best concentrations of SE are shown in Table 3. ELCAS and ELPHOS performed best at a concentration of 4000 mg/L while ELCHAS at 3000 mg/l. ELPHOS cost was twice that of ELCAS. The successful treatment of coffee factory‘s effluent using a 100 litre tank shows that this surface to volume ratio of 75 m2/m3 could be used as a scale-up ratio to fullscale treatment unit.

SCALE-UP OF A PROTOTYPE FOR COLOR REMOVAL In order to scale-up the laboratory reactor, instead of color, we can choose a compound present in the factory‘s effluent which will be called component (A), and the batch reactor below:

Feed Initial mole of Component (A) = NAO

Outlet Final mole of component (A) = NA

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Initial conversion of Component (A) = XAO=0 Initial molar Flow rate of Component (A) = FAO

Final Conversion=XA Final molar flow rate=FA

Reaction rate (-rA)

Figure 11: Batch reactor If a material balance is written around Figure 11, the equation around the reactor is: Feed in = Feed out + rate of reaction + rate of accumulation ………. …………………. 3 Feed in = Feed out = 0; since no feed is going in and no feed is coming out (batch treatment). So equation 3 becomes: (rate of accumulation: dNA/dt) = rate of reaction * volume of the reactor: (-rA)V ……….4 dNA/dt = (-rA)V ………………………………………………………...…………………5 NA = NAO(1-XA) Putting the value of NA into equation 5 We obtain: d[NAO(1-XA)]dt = (-rA)V …………………………………………………………………6 Since NAO is constant, its differentiation is equal to zero; so equation 6 yields: -[NAO dXA/dt] = rAV

………………………………………………………………….7

The following equation can be used to calculate for volume V Vt= [NAO]*[ dXA/rA] To do that, what one needs is to plot 1/rA versus XA; the area under the curve will be = V*t/NAO. Since in our experiment we did not study the reaction kinetic, we cannot get the plot 1/rA versus XA, which will be the ideal way to estimate the volume V. All things being equal we could use the volume Ratio. But since a direct ratio of Vp/Vlab = 150 is deemed too large, to get around the problem, we can use flow system of two to three reactors (Constant Stirrer Tank Reactors) in series. In the absence of the flow system, we will scale-up the reactor using the surface to volume ratio mentioned earlier.

i)

OPTION ONE Dimensions of tank before scaling-up Length 1 width 1 (cm) (cm) 65.0

45.0

height 1 (cm) 47.0

Volume 1 (cm3) 137475

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Scaling-up factor to construct a 15 m3 tank (0.65) x (0.47) x (0.45) X3 = 15 m3 X = (15/0.137)-1/3 = 4.78

iii)

Dimensions of new tank with 15 cubic meter capacity Length 2 width 2 height 2 (m) (m) (m) 3.11 2.15 2.25

iv)

Dimensions of each electrode 7/8 height width (cm) (cm) 42.0 30.0

Volume 2 (m3) 15.0

thickness (cm) 0.20

In the laboratory tank reactor, there were 6 cells each with three electrodes occupying a total surface area of 4.54 m2. To maintain a surface to volume ratio of 75 m2/m3 for a total volume of 15 m3, we need a total electrode surface area of 1125 m2. We calculate the total length of a cell in the laboratory reactor as 6.6 cm. Number of cells to fit in the length of the scaled-up reactor, after removing approximately 5 cm from either ends: (311.0 -10)/6.6 = 45.6 cells For this option, we will use 45 cells or 45 x 3 = 135 electrodes. CALCULATION OF CURRENT DENSITY Number of positive electrodes 45. Area of an electrode in the scaled-up tank is: 1125 m2/135/2 = 4.17 m2 At a required current density equals 1.20 Amps/m2, Current per electrode (one positive cell) is equal = 1.2 x 4.17 = 5.0 Amps Total reaction current in Amperes equals = 5.0 x 45 = 225 Amps CALCULATION OF POWER REQUIRED If the voltage per cell is equal to 40 volts, the minimum power required to run the color removal reaction is equal to: 40 x 225 = 9,000 W or 9.0 kW Assuming a Transformer and Rectifier with combined efficiency of 85% at 0.95 power factor (pf), the power rating is: 9.0 kW = 11.15 kW or 11.15 x (54/60) = 10.0 0.85 x 0.95 At US$0.25 the cost of a kWh in Kenya, the factory will be spending 10.0 x 0.25 = US$ 2.50 per day to treat its wastewater. Initial capital is may be high due cost of 15 m3 tank, hoists for the electrodes and their insulators. In addition there will be the need to adapt a welding transformer and a rectifier to provide a DC reactor current for only 54 minutes per day. OPTION TWO No scale-up of the laboratory tank, but setting 12 equal tanks side by side with a total volume of 1.2 m3

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Process:

1234-

Filling of tanks Color removal reaction Discharge to the stream

Preparation for the next batch Total time required for 1 single batch process

2 minutes 54 minutes 2 minutes

2 minutes 60 minutes

Volume per tank 0.10 m3 Number of tanks 12.0 Total volume 1.2 m3 3 Time required to treat 15 m wastewater from the factory per day: 15 (m3) 1.2 (m3/batch)

x

60 (minutes) batch

x

(1) (hr) (60 minutes)

= 12.5 hours

Active surface area of an electrode = 0.126 m2. Since there are twelve (12) positives electrodes in a tank, the total surface is 0.126 x 12 = 1.51 m2. For a current density of 1.2 Amps/m2, the total current going into a tank is 1.2 x 1.51 = 1.81 Amps. There are twelve tanks in this option. The amount of current required is 1.81 x12 = 21.74 Amps per batch. Assuming we maintain the same voltage of 40 volts, the power rating here will be: 40 x 21.74 = 869.6 W or 0.869 kW per batch. Assuming a Transformer and Rectifier with combined efficiency of 85% at 0.95% power factor (pf), the power rating will be: 0.869/(0.85 x 0.95) = 1.08 kW Knowing that there will be 12.5 batches per day of 60 minutes each, the total power consumed will be: 1.08 x 12.5 = 13.5 kW or 13.5 x (60/60) = 13.5 kWh Total cost of power consumed = 13.5 x 0.25 = US$ 3.36 per day to run the color removal reactors. In each of the two options presented above, the costs of control equipment (pumping, draining, logic control) as well as the cost of labor has not been included. It appears that the 1st option is the less expensive of the two.

CONCLUSION AND RECOMMENDATION This project has shown that rock phosphate can be a good substitute for wood ash during electrochemical colour removal. ELPHOS can also yield an effluent whose quality can be considered for re-use in the factory for cleaning purposes. However, the high cost of this process of $6,816/year may be a deterrent for an otherwise effective method that can completely remove colour and substancially reduce COD and BOD. Using these results, we have showed that it is possible to determine the reaction time and volume of an industrialscale reactor required to treat effluent from a coffee factory effluent. However additional studies are necessary to verify these results and further refine parameters for industrial application. It is therefore suggested that more studies be carried out to make this method more economical by, for instance, increasing the electrode surface area or by recycling the treated water.

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ACKNOWLEDGMENT The authors wish to acknowledge the assistance of Moi University, Kenya for funding this project and Ms Abigael Nekessa and Mr. Thuita Moses, graduate students at Moi University, Department of Soil Science, School of Agriculture and Biotechnology for providing the necessary information on phosphate rock used in this experiment.

REFERENCES Ahonen, H. (2001). A Review of General Information on Electrochemical Process Wastewater Treatment. Finland Laboratories 7p. Arudel, J. (2000). Sewage and Effluent Treatment. Wiley and Sons, Great Britain, 45-90. ASTM. (1983). Standard Practice for Coagulation-Flocculation Jar Test of Water. Annual Book of ASTM Standards. 1101. Coste, R. (1992). Coffee - The plant and the product. MacMillan Press, London. Etiégni, L. & Campbell, A. G. (1991). Physical and Chemical Characteristics of Wood Ash Bioresource Technology, Vol. 37, 173-178. Etiégni, L., Orori, B. O. & Rajab, M. S. (2004). An Electro-coagulation Method of Color Removal in Wastewater or Water with Low Power Consumption. International Patent (PCT/ KE2005/000012). Ikeheta, Z. P. & Buchanan Q. S. (2000). Decolorization of pulp mill effluents with immobilized lignin and manganese peroxidase from Phanerochaete chrysosporium. Environ. Technol., 19, 521-528. Kashefialasl, M., Khosravi, M., Marandi, R. & Seyyedi, K. (2006). Treatment of dye solution containing colored index acid yellow 36 by electrocoagulation using iron electrodes. Int. J. Environ. Sci. Tech., Vol. 2(4), 365-371. Koparal, A. S. & Gütveren, Ö. B. Ü. (2002). Removal of nitrate fromwater by electroreduction and electrocoagulation, J. Hazard. Mater, B 89, 83-94. Muleta, D. (2007). Microbial inputs in coffee (Coffea arabica L.) production systems, South Western Ethiopia. Implications for promotion of Biofertilizers and Biocontrol agents. Doctoral thesis, Swedish university of Agricultural Sciences. Orori, O. B., Etiégni, L., Rajab, M. S., Situma, L. M. & Ofosu-Asiedu, K. (2005). Decolorization of a pulp and paper mill effluent in Webuye Kenya by a combination of electrochemical and coagulation methods. Pulp and Paper, Canada, Vol. 106, No. 3, T50T55, 21-26. Orori, O. B. (2003). Colour removal from wastewater of a Pulp and Paper Mill in Kenya, by a combination of electrochemical and coagulation method. Mphil. Thesis, Department of Wood Science & Technology, Moi University, Eldoret, Kenya, 146. Othman, F., Sohaili, J., Ni‘am, F. M. & Fauzia, Z. (2006). Enhancing suspended solids removal from wastewater using Fe electrodes. Malaysian Journal of Civil Engineering, 18(2), 139-148. Prasad, D. Y. & Joyce, T. W. (1991). Colour removal from Kraft bleach plant effluents by Trichoderma sp. TAPPI J., Vol. 65, No. 1, 165-169.

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Springer, A. M., Vincent, C. H. & Timothy, S. J. (1995). Electrochemical Removal of Colour and Toxicity from Bleached Kraft Effluents. TAPPI J., Vol. 78, No. 12, 85-91. UNEP. (1981). Environmental Management in the Pulp and Paper Industry, Vol. 1, 234. UNESCO/WHO. (1978). Water Quality Surveys - A guide for Collection and Interpretation of Water Quality Data. United Kingdom, 350. Van Kauwenbergh, S. J. ( 1991). Overview of phosphate deposits in Eastern and Southern Africa. Fert. Res., 30, 127-150. Wei-Lung, C., Chih-Ta, W. & Kai-Yu, H. (2009). Effect of operating parameters on indium (III) ion removal by iron electrocoagulation and evaluation of specific energy consumption. Journal of Hazardous Materials, In press doi:10.1016.

In: Fluid Waste Disposal Editor: Kay W. Canton, pp. 299-314

ISBN: 978-1-60741-915-0 © 2010 Nova Science Publishers, Inc.

Chapter 14

WATER AS A SCARCE RESOURCE: POTENTIAL FOR FUTURE CONFLICTS M. A. Babu* Department of Environment, Islamic University in Uganda; P.O .Box 2555, Mbale- Uganda.

ABSTRACT The major aim of this paper is to review the major problems of water resources in the developing countries. It is based on problems related to population growth and pollution and how these are more likely to lead to future conflicts. We know that fresh water is only 3 % of the total global water and 78% of this is in glaciers. This makes it a scarce and precious resource which must be sustainably managed. The paper also analyses some of the already existing and potential conflicts based on water resources. It reviews the potential threats to Ugandan water resources and problems which are most likely to occur as a result of these threats. Factors hindering treatment of wastewater as a remedy to pollution in developing countries have also been discussed. The methodology used in this paper is based on literature review of the most current issues that affect water resources world-wide. The review is limited to scientific facts and no political factors affecting water resources have been included. It has been found that although Uganda is endowed with 66km2/year of renewable water resources, population increase, deforestation, degradation of wetlands and pollution are major threats to its water resources. Problems associated with water quality and quantities are more likely to result into internal conflicts which are bound to spread beyond Ugandan borders.

Keywords: Water scarcity, deforestation, population, wetlands, pollution, conflicts.

*

Corresponding author: Phone: +256 45 33502, Fax: +256 45 34452, Email: [email protected]

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INTRODUCTION Oil is an invaluable resource to mankind; it makes us fly, drive, generate power, till the land and even drive military hardware. It is a potential threat to the western world and source of conflict in the world politics. It is believed to be the major destabilizing factor in the Middle East (Darwish, 1994). It is a luxury that mankind cannot afford to live without. On the other hand, water is part of us - the human body is composed of 70- 95% water, the food we eat is water in a different form. It is an essential component that drives the ecosystem and the food chain. It is a resource with inherent values that cannot be comparable and measurable. Imagine if water were to be mined like oil, would we live and survive? Imagine New York City without water for 24 hours! Water is a basic need; it is not a luxury that mankind can live without. Water is becoming an increasingly vital resource and is thought to over take oil as the potential cause of conflict in the Middle East. President Anwar Sadat once said Egypt will never go to war except when its water resources are threatened. According to UN- reports, the population in the Middle East is increasing rapidly (Table 1.0) and it is predicted that this will exert more pressure on the already existing problems associated with water resources. In the UN report, 18 countries will be on the list of water scarce countries by 2025. These include: Algeria, Israel/Palestine, Qatar, Saudi Arabia, Somalia, Tunisia, United Arab Emirates, Yemen, Egypt, Ethiopia, Iran, Morocco, Oman and Syria. Water scarcity can be described as when country has less than 1,700 m3 per capita, it is said to be experiencing water stress, while less than 1000 m3 is regarded as water shortage. This list includes most countries that are already in the volatile Middle East region hence the possibility of conflict cannot be ruled out. Past experience of the 1960s is testimony to the conflicts. Cross border raids on water schemes' between Israel, Syria and Jordan culminated into the six day war in 1967 (Darnish, 1994). Although natural water resources can be replenished through the hydrological cycle, then one might urge that the natural cycle will take care of the situation. The average renewal rate for rivers are about 18 days while large lakes and deep aquifer can take up to thousand of years. The Nubian aquifer in North Africa- known as the world's oldest aquifer were thought to be filled in past geological years. When depleted, it is not known how long it would take to recharge (Darnish, 1994). Climate change, pollution and the demand to feed large populations makes the problem more complex. It is predicted that the main conflicts in Africa in the next 25 years could be over water. It is thought that countries will fight each other to have access to this precious and scarce resource (Russel, 2007). According to the UNDP report as cited by Russel, (2007), 12 African countries will join the list of water scarce countries. This means that a person will have less than 1000 m3 of water per year. Water requirement for domestic use depends on lifestyle and availability. For instance, about 400, 200 and 10 - 20 litres per person per day are consumed in North America, Europe and Sub-Saharan Africa respectively (World Water Council, 2000). As can seen, the current consumption in Sub-Saharan Africa is the lowest and it is thought to worsen in the next 50 years. The Nile, Niger, Volta and Zambezi basins are believed to be major areas of conflicts. Other potential water war areas in Southern Africa will involve Botswana, Namibia and Angola. There are already tensions in these regions.

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Table 1.0. Prediction of population (in Millions) of some selected Middle East countries. Country Israel Palestine Libya Egypt

1990 4.7 4.5 58

2025 8 7 12.9 101

(Source: Darnish, 1994)

Figure 1. Possible regions of water wars in Africa (UNDP, 2007).

According to Lester Brown (as cited by Russel, 2007), the head of environmental research institute Worldwatch, the combined population of Ethiopia, Sudan and Egypt – will rise from the current 150 million to 340 million in 2050. This will result intense competition for increasingly limited water resources. "There is already little water left when the Nile reaches the sea," he says. Egypt is ready to use force to protect its vital water resources. It is much more concerned with dams that might be constructed in the Ethiopian highlands which are thought to affect the flow of the Nile. Egypt is not in compromising position as regards this. In 1989, the Ethiopian ambassador in Egypt was summoned to explain the presence of Israel hydrologists around the areas of the Blue Nile. During the same period, the Egyptian parliamentarians declared their willingness to back their government in taking military action against Ethiopia if it becomes necessary (Darwish, 1994). In the Niger basin, Senegal and Mauritania have already fought two short wars in 1987 & 1989 over the Senegal River. The cause of this was that Mauritanian tribesmen searching for vegetation crossed to the other bank, violating Senegalese sovereignty (Darwish, 1994).

UGANDAN SCENARIO Uganda is naturally endowed with water resources; renewable water resources are estimated to be 66km2/year which corresponds to 2800m3/person/year. Open water surface covers 15% of the total land cover (Figure 2). Uganda also receives annual average rainfall of 900-2000mm of which 7-20% of this recharges the underground water.

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Figure 2. Major Lakes and Rivers of Uganda (WWAP, 2006).

The recharge to ground water is high as compared to the current abstraction volumes hence there is no over exploitation at current situation. However, it should be noted that Uganda has 106 towns of which 56 towns have piped water supply. Of the 56 towns, 15 large towns are mainly supplied with lake or river water. The remaining 41 small towns depend on high yield boreholes (WWAP, 2006). It should be borne in mind that these towns are rapidly expanding and the demand for water may result into tapping more of underground water sources. At same time, there are also 69 other small towns that do not have piped water supply. If we are to meet the target for the Millennium development goals by supplying water to the towns, the underground sources may be strained.

Major Threats to Water Sources Uganda has not yet fitted into the larger picture of water wars. Whether it will feature or not can be established through, analysing major factors that may drag it into action. In this context, these are limited to only scientific and not political factors. In our own opinion, threats to water sources may become future points of conflict as time goes by. The following anthropogenic activities may be considered as threats to Ugandan water resources:

(a) Deforestation It is well documented that areas of high altitude with dense rain forest covers receive high precipitation (WWAP, 2006).The forests provide moisture in the hydrological cycle required for rainfall formation. At the same time, vegetation cover traps and holds rain water; realising it slowly hence recharging rivers, streams and underground wells throughout the year.

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2000000 1800000

Forest cover ( ha)

1600000 1400000 1200000 1000000 800000 600000 400000 200000 0 2005

2010

2015

2020

2025

2030

2035

Time (years)

Figure 3. Forest cover predicted to be lost in the next 30 years.

Of recent, there has been enormous pressure on forest systems. According to Environmental profile of Uganda (2005), Uganda had total forest cover of 3,627,000 ha (prior to 1970) but this has been reduced by 50% from 1971-1987. Again from 1990-2005, 26.3% of the remaining forest covers has been lost through deforestation. The country profile also reports annual deforestation rates of 2.2% per year for a period from 2000-2005. Taking a conservative approach and assuming that by 2007 the forest cover is 50% of 3,627,000 ha and the annual deforestation rates of 2.2% per year is maintained, the total forest cover will be less than 200,000 ha in the next 100 years (Figure 3). Forests have been cleared for settlement, agriculture as well as development. Notable cases include the recent proposal of give away of Mabira forest. Already, some of the forests on the Islands of L.Victoria have been given away for palm tree growing. The forests around Mt. Elgon region are already generating conflicts between the local community and Uganda Wild Life Authority. The net result of forest loss is reduction in precipitation thus many river systems running dry. The water holding capacity of vegetation will be eliminated hence water will be flushed through rivers at faster rate; leaving them dry especially in the dry seasons. The highlands of Uganda are big producers of fruits and vegetables like cabbages, tomatoes and onions, as well as non vegetable crops like Irish potatoes bananas, maize and coffee. Loss of soil fertility due to soil erosion coupled with increased flushing of rivers will affect the livelihoods of many people. Water availability, proximity, quality and quantity have been strongly linked to poverty (UN, 2007). Siltation of rivers and poor water quality will be come common, diseases and poverty will be alleviated and this may be a flash point of conflict within the communities. The export of virtual water and nutrients in crops from areas already constrained by water shortage and loss of nutrients (due to erosion) will further make producing areas poorer. Unfortunately, the water and nutrients in crops are exported to urban areas where they are flushed down the drain without a possibility of recycling. It also known that the highland forests act as catchments for the wetlands down stream, a case in point being the Manafa – Doho drainage system. The extent of flooding in the plains

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down stream will be limited to smaller areas that can be sustained by the little water from the mountains. Competition for the limited wet areas available will cause tension in the rice growing communities. Of recent, violence linked to possession of wetlands has been reported in this area. It is also thought that rifts may develop between the communities up and down streams as they scramble for little water available. Worthy noting, there be will silent conflicts between man and biodiversity for the limited resources.

(b) Wetlands Uganda has a total wetland area of 30,000km2. Like forests, wetland degradation is on the increase. For instance, it is estimated that 45 % of the Nakivubo wetland has been modified or reclaimed (Emerton et al, 1999). The Nakivubo wetlands as well as those in Eastern and other parts of Uganda are under pressure. Key activities that degrade wetlands include agriculture, development, settlement and solid waste disposal. It has also been found that most of the wetlands of the valley bottoms have been converted to agricultural use. This has led to change in micro climate and even lowering of water table as seen in Kabale and Bushenyi (WWAP, 2006). There are 3 hydrological functions of wetlands which are important in maintaining the hydrological balance. First, wetlands act as holding water basins slowly releasing water into rivers and streams thus ensuring continuous flow throughout the year. Secondly, the holding basins also provide more time for seepage and recharge of underground water and thirdly, wetlands play a vital role in providing water directly to its beneficiaries. It is estimated that wetlands directly provide water to 5 million people in Uganda (WSSP, 2001).Hence, off setting the hydrological balance will lead to drying up of most rivers, streams and wells. The quantity of underground water sources is expected to decline. Loss of wetlands may mean loss of many livelihoods. Many wetlands are used for fishing, grazing, agriculture and for materials used in shelter. Effects of wetland loss are likely to be felt by the poor rural communities who largely depend on them. Loss of livelihood may become source to conflicts.

(c) Pollution This is a major problem mainly facing urban areas. L.Victoria is of major interest since it is a vital component in the Nile basin, being a resource shared by many riparian countries. It is the second largest fresh water lake in the world with an area of 69,000km2. The lake is increasingly receiving pollution from un-treated sewage as well as industrial effluent. Odada et al., (2004) report that the number of people without sewers in urban areas of L. Victoria region is high and yet the population growth is over 5- 10%. This raises concern in regard to the lake quality. According to UNEP-SEO report (2004-2005), the lake is experiencing severe impact on microbial contamination, eutrophication and suspended solids. It is also seen that the Congo basin which is within the catchments areas of L.Victoria is under severe eutrophication (Table 2.0).In fact, it is anticipated that the cost of water treatment for Kampala city is bound to rise due to increased levels of phytoplankton in the lake. The city draws its raw water from the lake yet the demand for water supply is fast tracking the rise in population.

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Table 2.0. Water pollution in selected water bodies (UNEP, SEO Report, 2004/2005).

Most industries located in Kampala and probably elsewhere in Uganda do not have effluent treatment systems but drain their wastes into the lakes, rivers and environment. It has been found that industries release 1,045kg/d BOD5, 96kg/d of nitrogen and 105kg/d of phosphorous into the lake (WWAP, 2006). The BOD exerts oxygen demand on the water affecting fish and other aquatic organisms. Nitrogen and phosphorous cause algal blooms, which may cause skin irritation, production of toxins as well as increasing oxygen depletion in the lake. Cholera cases have become common in Kampala and cases of dysentery have increased from 2300 in 1999 to 8300 in 2002 (WWAP, 2006). Apart from the industrial effluent into L.Victoria, there are a number of flower farms being established on the shore line of the lake. Flower farms are known to extensively use fertilizers and pesticides. If this is left unchecked, water from the lake may become unfit for consumption or else we pay the heavy price of treating pesticides. As for the rural areas, small towns are rapidly cropping up and more pit latrines are being sunk. Underground water will be contaminated and incidences of water borne diseases will be high. Much as Uganda is endowed with water resources, lack of mitigation to pollution will result in loss of this precious resource.

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(d) Population growth Uganda like many developing countries is experiencing rapid population growth. With the invention and perfection of the Haber process in the 1960‘s, population and food production has increased synonymously. The Haber process has greatly improved the manufacture of nitrogen based fertilizers. Nitrogen in the fertilizers is transformed to proteins in the food chain and upon consumption followed by excretion, large amounts of nitrogenous products are released into the environment (Mulder, 2003). This is known to pollute the receiving water bodies. Nitrogenous wastes, organic matter and pathogens found in wastewater have affects on public health and harmful ecological impacts on the environment (Gijzen and Mulder, 2001). The growing concern of nitrogen, organic and pathogen pollution therefore calls for the need of wastewater treatment before discharge into natural watercourses. At present, there is hardly any infrastructure for effective treatment of sewage in developing countries. Municipal sewerage system coverage and the extent of domestic and industrial wastewater treatment are inadequate. Treatment level is insufficient in most urban situations (Gijzen et al., 2004). For example in Latin America, only 14% of collected sewage receives treatment (WHO/UNICEF, 2000). Even when the facilities exist, poor maintenance and operation results in failure of treatment processes causing pollution of the receiving surface waters (Gijzen et al., 2004). In Sub-Saharan Africa, 42% of the population (as per 2002) lacked water supply and 64% lacked basic sanitation (WHO-UNICEF, 2006). Reduction of these proportions will definitely have environmental effects. Even if these percentages are spread to both urban and rural communities, the ultimate end of the waste is the environment. This poses a challenge to most governments. The major hindrance to sewerage coverage and treatment systems in most developing countries is the weak economies. Priority in these states is given to security, health and education. For instance in Uganda, the 1998/1999 budget for sanitation dropped from 46%18% while that of education increased from 14-47% (WSP, 2004). Also the cost of conventional wastewater treatment infrastructure is prohibitive for the majority of these developing countries, Uganda being inclusive (Gijzen et al., 2004). The implementation of conventional wastewater collection and treatment in developing countries to attain EU standards is therefore unrealistic, except maybe in densely populated urban centres where the average income is much higher. However, this should not be generalised as most urban areas are facing problems of their own. Population growth in these areas is seen to out grow the existing wastewater treatment infrastructure (Gijzen and Khonker, 1997; Yu et al, 1997). Weak economies, corruption, increased demand for urban land and unplanned development seems to make the expansion of the existing infrastructure nearly impossible (WWAP, 2006). Although the millennium development goals emphasize strong water and sanitation component- which is good, there is a possibility that this will compound the problem of water pollution further if not well handled. Goal seven for instance proposes reduction of half of the proportion of people without access to safe drinking water and basic sanitation by 2015 (UN, 2007). It is estimated that 2.6 billion people lack access to sanitation (WHO-UNICEF, 2006). If this goal is to be achieved; there will be increased generation of wastewater. Mara et al., (1992) estimate that 80-90% of water consumed is converted to wastewater. Increased production of wastewater will put more burdens on the already strained waste management.

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From this point of view, it is realistic that governments should think of strategies of solving problems of water pollution if they are to achieve the millennium development goals and avoid future conflicts. They should make use of low cost treatment technologies through best approach to improve effluent quality. Much as there is evidence of success of conventional approach, the concept still needs to be reconsidered from sustainability point of view (Gijzen et al., 2004). As such, efforts on improving nutrient removal using natural wastewater treatment technologies like wastewater stabilization ponds and wetlands that are widely applied in developing countries become paramount. This will entail protection of the environment from pollution resulting from wastewater discharge. Governments should reduce the pressure on the already existing treatment infrastructure by avoiding the centralized approach of sewerage collection. It is recommended that they delocalize collection systems to many smaller but effective treatment systems. It is expected that with the growing population coupled with destruction of forests, wetlands and increased pollution, the quality and quantity of available water will reduce. This is most likely to spark conflicts in regions which are already facing water scarcity.

What Do We Learn? If issues of deforestation, wetland degradation, and pollution and population growth are not addressed; Uganda is bound to face problems of water quality and quantity. There will be more internal conflicts based on water resources. Seasonal rivers will dry up, tribal conflicts will be elevated and pastoralists will be forced to move to other areas in search of water and pasture. Already the Balalo and Basongora pastoralists in Uganda have generated a lot of ethnic sentiments. Internal conflicts in Uganda may spread beyond its borders. Sudan and Egypt will be definitely affected and Egypt may be forced to extend its long arm to Uganda. Cost of water treatment in Uganda will increase and the less privileged who cannot afford will be vulnerable to water borne diseases and exposure to pollution. Poverty eradication may be a dream. Water is a scarce and precious resource which must be managed and used sustainable. If not well, managed water is likely to cause conflicts world wide.

ACKNOWLEDGMENTS Am grateful to Dr. N. Sarah for sparing her time to read and review this work. Her contributions have added invaluable knowledge to this paper.

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REFERENCES Darwish, A. (2004). A Lecture on Environment and Quality of Life, June 1994. Geneva conference. Emerton, L., Iyango, L., Luwum, P. & Malinga, A. (1999). The Present Value of Nakivubo Urban Wetland, Uganda. National Wetlands Conservation & Management Program/ IUCN. Environmental profile (2005). Forest cover of Uganda. www.rainforests.mongabay.com Gijzen, H. J, Bos, J. J., Hilderink, H. B. M., Moussa, M., Niessen, L. W. & de Ruyter van Steveninck, E. D. (2004). Quick scan health benefits and costs of water supply and sanitation. Netherlands Environmental Assessment Agency. National Institute for Public Health and the Environment – (MNP-RIVM), The Netherlands Gijzen, H. J & Mulder, A. (2001). The global nitrogen cycle out of balance. Water, 21, Aug 2001, 38-40. Gijzen, H. J. & Khondker, M. (1997). An overview of ecology, physiology, cultivation and application of duckweed, Literature review. Report of Duckweed Research project. Dhaka, Bangladesh. Mara, D. D., Alabster, G. P., Pearson, H. W & Mills, S. W. (1992). Waste stabilization ponds, a design manual for Eastern Africa, Lagoon Technology International Leeds, England. Mulder, A. (2003). The quest for sustainable nitrogen removal technologies. Wat Sci. Tech., 48(1), 67-75. Odada, E. O., Olago, D. O., Kulindwa, K., Ntiba, M. & Wandiga, S. (2004). Mitigation of environmental problems in L. Victoria, East Africa: Casual chain and policy option analyses. Ambio, Vol. 33(1-2), Feb. 2004. Russell, S. (2007). Africa’s Potential Wars, BBC online. www.bbc.com UN (2007). The Millennium Developments Goals Report 2007, New York. WHO-UNICEF (2006). Meeting the MDG drinking water and sanitation target: the urban and rural challenge of the decade, WHO press, Geneva. World Health Organisation / UNICEF. (2000): Global Water Supply and Sanitation. Assessment 2000 Report. Geneva, World Health Organisation. World Water Council (2000). World Water Vision, Making water everybody’s business. The Use of Water Today, The Hague, Netherlands WSP, (2004). Strengthening Budget Mechanisms for Sanitation in Uganda. WSSCC (2004). Resource packs on the water and sanitation Millennium development Goals. Water supply and sanitation collaborative council, Geneva. WSSP, (2001). Wetland Sector Strategic Plan 2001-2010. Wetlands Inspection Division, Ministry of Water, Lands & Environment – Uganda. WWAP (2006). Uganda National Water Development Report, UN report. Yu, H., Tay, J. & Wilson, F. (1997). A sustainable municipal wastewater treatment process for tropical and subtropical regions in developing countries. Wat. Sci. Tech., 35(9), 191-198.

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ISBN: 978-1-60741-915-0 © 2010 Nova Science Publishers, Inc.

Chapter 15

RECYCLING WASTEWATER AFTER HEMODIALYSIS: AN ENVIRONMENTAL AND COST BENEFITS ANALYSIS FOR ALTERNATIVE WATER SOURCES IN ARID REGIONS Faissal Tarrass*1, Meryem Benjelloun 1, and Omar Benjelloun 2 1

2

Hassani Hospital Center, Nador, Morocco Hospital Universitario Central de Asturias, Oviedo, Spain

ABSTRACT Water is a vital aspect of hemodialysis. During the procedure, large volumes of water are used to prepare dialysate and to clean and reprocess machines. This paper evaluates the technical and economical feasibility of recycling hemodialysis wastewater for irrigation uses, such as watering gardens and landscape plantings. Water characteristics, possible recycling methods, and the production costs of treated water are discussed in terms of the quality of the generated wastewater. A cost-benefit analysis is also performed through comparison of intended cost with that of seawater desalination, which is widely used in irrigation.

Key words: Hemodialysis, Environment, Wastewater, Water quality, Recycling, Membrane technology

INTRODUCTION Water is essential to all known forms of life, but this resource is under threat [1]. Growing national, regional, and seasonal water scarcities in much of the world pose severe challenges for national governments, international development, and environmental policies [2-5]. In this context, alternative water sources such as wastewater recycling offer a partial solution by creating fresh water for industry and agriculture [6].

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Hemodialysis, a method for replacing renal function in patients suffering from renal failure by the removal of excess water and wastes, requires a large volume of water. Water is used in long-term dialysis facilities to prepare dialysate and to rinse and reprocess dialysis membranes and machines [7,8]. Assessing the recycling potential of hemodialysis wastewater must give consideration to both the environmental and economic aspects. This study was undertaken to analyze hemodialysis wastewater at a Moroccan dialysis facility in order to examine the feasibility of advanced wastewater treatment for agricultural uses such as watering gardens and landscape plantings.

WASTEWATER GENERATION IN HEMODIALYSIS Water is a vital aspect of hemodialysis. During hemodialysis, assuming a dialysate flow rate of 500 mL/min, a patient is exposed to 120 liters of purified water during a typical 4-hour dialysis session. The yearly consumption of water for a single-pass dialysis system operating 12 hours per day and 6 days a week is estimated to be 112 m3 [8], without considering water that is rejected during treatment by the carbon filters and reverse osmosis membranes prior to dialysis use [8]. In Morocco, there are 135 dialysis centers, offering 1589 hemodialysis stations. The number of patients under treatment is estimated to be 5737 patients on September 2007, with a number of treatments of 8 x 104 per year approximately [9]. A simple calculation shows that the total consumption of water by hemodialysis facilities exceeds 50 million US gallons per year (1 m3 = 264 US gallons). This figure is excessive for a country experiencing drought [10]. Minimizing this water expenditure is of importance to Morocco and other water-poor countries in which scarcity of water sources represents a serious impediment for long-term development [2-6].

IMPACTS OF RECYCLING Hemodialysis wastewater can enter municipal and natural water systems via residential or commercial discharges, including hospital effluent. There is a lack of data about the possible direct and indirect impact of hemodialysis wastewater discharges on the environment. However, benefits of recycling may result in reduced discharge of wastewater into natural water bodies and the potential water savings for hospitals. In terms of economic impacts, it is known that integrating wastewater treatment in agriculture can bring benefits such as partial cost recovery [1,2]. For hospitals, recycling wastewater may provide a purchase price reduction [11]. In this study, we analyze the potential economic benefit of using recycled hemodialysis wastewater for irrigating the hospital grounds.

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METHODOLOGY Wastewater Sampling Water samples were obtained from a single dialysis facility. Using sterile 500-mL bottles, wastewater was collected from the outflow pipe that drains all hemodialysis sewage (including waste dialysate and water rejected during treatment by the carbon filters and reverse osmosis membranes) directly into the municipal sewage line. All samples were placed in a closed cooler during transit to the laboratory.

Wastewater Chemical Analyses and Physical Characteristics Wastewater samples were analyzed in an accredited laboratory for biochemical oxygen demand, total Kjeldahl nitrogen, phosphorus, chloride, and sulfate using spectrophotometry (Hache Lange Company, Noisy le Grand, France) [12]. Biochemical oxygen demand is a measure of the amount of oxygen utilized in the biochemical oxidation of organic matter present in water. Temperature, pH, and conductivity analyses were also performed.

Wastewater Microbiological Analyses Samples for bacteriological testing were processed within 1 to 2 hours. Samples were cultured using the membrane filtration technique. In brief, membrane filters were placed aseptically on trypticase soy agar and incubated at 36ºC for 48 hours [12,13]. Total viable colony counts were documented and isolates identified using standard microbial techniques.

Wastewater Quality Criteria The suitibility of hemodialysis wastewater use in agriculture was evaluated through the comparison of its characteristics with the Food and Agriculture Organization of the United Nations (FAO) and the World Health Organization (WHO) standards for wastewater use for agricultural applications [14,15]. The optimum procedure for treatment to reach standards is discussed according to the quality of the generated wastewater.

Cost Estimation Analyses CapdetWorks, a preliminary design and costing program available from (Hydromantis, Inc, Ontario, Canada) was used for the design and cost estimation of wastewater treatment plants. The design and cost estimation of reverse osmosis plants was modeled with WTCost, which is a preliminary design and costing program for processing plants developed with the support of the US Bureau of Reclamation [16]. Both tools incorporate water quantity (average daily, maximum, and minimum flow), water quality (influent wastewater characteristics and

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desired effluent quality), and land requirements. The cost report produced by the two programs includes equipment, operation and maintenance cost (materials and supplies, energy, and labor). A cost benefits analysis was also performed in which the cost of water treatment was compared with that of seawater desalination to produce water of equivalent quality. Data for cost estimates of water desalination for agricultural applications were based on the proceedings of the FAO expert consultation on water desalination for agriculture [17].

RESULTS Hemodialysis Wastewater Composition and Pollution Risk Chemical characteristics of the wastewater are displayed in Table 1, as well as information on other residual wastewater that is commonly recycled, and the FAO/WHO standards for wastewater use in agriculture [14,15,18-21].

Table 1. Comparison of hemodialysis wastewater composition with other commonly recycled waters, wastewaters and quality standards for agriculture Parameter

Hemodialysis wastewater (this study)

Temperature (°C) 30 PH 7.84 Conductivity (µs/cm) 13200 Biochemical oxygen 24.2 demand (mg/l) Ammonia nitrogen (mg/l) Kjeldahl nitrogen (mg/l) 29 Chloride (mg/l) 289 Sulfate (mg/l) 80.4 Phosphorus (mg/l) 54 Bacterial count 450 (CFU/ml) TDS (mg/l)

Seawater Municipal [20,21] Morocco United (LYDEC) States [19] [18]

Industry

22 – 30 7.6 – 7.9 50000 3.5 – 9.6

30 5.5 – 8.5 7.3 – 7.7 2520 500 110 – 400

30 6.5 – 8.5 2700 2190 50 – 600 204

0.2 – 1.1

12 – 50

18900 2649

34483

12 – 50

150 – 200 20 – 85 120 4 – 15 22 x 108 500

Morocco United [19] States [18]

FAO/WHO standards for irrigation water [14,15] 6 – 8.5 300 – 700 5 – 45

39.5

0–5

397 270 11.2

0 – 30 0 – 20 0–2 2x104– 10x104

250 – 850 50 – 600 1406

< 450

Abbreviations: LYDEC, Lyonnaise des eaux de Casablanca; FAO, Food and Agriculture Organization of the United Nations; WHO, World Health Organisation; CFU, Colony-forming unit; BOD, Biochemical oxygen demand = quantity of oxygen utilized in the biochemical oxidation of organic matter present in water; TDS, Total Dissolved Solids = solids that either float on the surface or are suspended in water

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These results show that apart from an increased but expected conductivity value, biochemical oxygen demand, Kjedahl nitrogen, chloride sulfate, and phosphorus concentrations did not exceed the FAO standards, with the exception of the conductivity value. Bacterial count of the wastewater showed 450 colony-forming units/mL, but coliform organisms (specifically Escherichia coli species) were undetectable. The primary challenge for use of hemodialysis wastewater can be its high conductivity. By contrast, concentrations of dissolved organic substances were under applicable emission standards for discharges [18,19], and the bacterial count was under WHO standards for wastewater use in agriculture [14].

Treatment Options for Recycling Wastewater Due to the high conductivity of hemodialysis wastewater, the use of membrane technology could be suitable for the treatment of such wastewater [22,23]. This technology has proven to be efficient and cost saving in comparison with other processes [24,25]. Computed simulations for cost estimation were performed for two possible membrane filtration processes, nanofiltration and reverse osmosis. Technical parameters of both processes are shown in Table 2.

Table 2. Technical characteristics of the nanofiltration and reverse osmosis systems for wastewater recycling Parameter Pressure (psi) (bar) Flow (m3/h) (US Gallon /day) Recovery (%) Power (System + Pumps) (kWh) Membrane life (year) Price (US$)

Nanofiltration

Reverse osmosis

100 – 150 7 – 10.5

100 – 150 7 – 10.5

1 – 1.2 6500 – 7500 30 2.2 3 3500

0.9 – 1.0 5500 – 6500 30 3 4 4500

Note: Values shown are for a thin film composite (TFC) membrane type. Abbreviations: psi, per square inch; kWh, Kilowatt hour

Cost Benefits Total costs for the two treatment techniques, including capital equipment, operating, and maintenance costs, are presented in Table 3. Costs were calculated based on current membrane and equipment prices. Energy, labor, and maintenance were calculated based on the average Moroccan prices for energy and labor. Membrane life was set at 3 and 4 years for nanofiltration and reverse osmosis respectively.

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The cost for treating hemodialysis wastewater to achieve a quality that is suitable for irrigation using nanofiltration and reverse osmosis is 0.70 US$/m3 and 0.74 US$/m3, respectively.

Table 3. Estimation of costs for wastewater recycling calculated on 288 working days per year, and 20 working hours per day Item Equipment (US$) Operating (O&M) Energy/working hour (US$) Labor/year (US$) Membrane replacement (US$) Cleaning chemicals/week (US$) Repair and maintenance/year (US$) Total costs/year (US$) Cost of production (US$/m3)

Nanofiltration 16500

Reverse osmosis 18000

0.0946* 14400 1050 0.90 175 3253 0.70

0.129* 15600 900 – 220 3423 0.74

Note: The plant capital was calculated on current membrane and equipment prices. The energy costs were based on Moroccan prices 26 (*0.043US$/kWh), labor costs on average Moroccan costs and the membrane replacement on an average 3 and 4 years replacement respectively.

For comparaison, the costs associated with various techniques for desalination of seawater for agricultural use are listed in table 4. With the exception of the latest reserve osmosis technologies, costs are in excess of 1.15 US$/m3 Given the average cost of 1 US$/m3 for seawater desalination [17], this could result in cost savings (or benefit) of 20 to 30% in comparison to desalination of seawater (Table 4).

Table 4. Comparison of costs of hemodialysis wastewater treatment versus desalination Parameter

Sea water desalination for agriculture use [17] Multistage flash (US$ /m3)

HD wastewater treatment for agriculture use

Capital cost Energy Fuel Electricity

0.301

Multipleeffect distillation (US$/m3) 0.520

Vapour compression (US$/m3)

Reverse osmosis (US$ /m3)

Nanofiltration Reverse (US$/m3) osmosis (US$ /m3)

0.548

0. 301

0.358

0.39

0.62 0.18 – 0.19

0.55 0.08 – 0.09

0 0.60 – 0.66

0 0.26 – 0.32

0.118

0.161

Labor Chemicals Membrane replacement Maintenance Payback costs Total costs

0.038 – 0.043 0.036 – 0.048 0.064 – 0.096 0.021 – 0.097 0.312 0.033 – 0.047 0.028 – 0.038 0.022 – 0.038 0.019 – 0.056 0.009 0 0 0 0.001 – 0.043 0.227

0.338

0.021 – 0.038 0.021 – 0.038 0.019 – 0.032 0.021 – 0.038 0.037 0.40 – 0.42 0.40 – 0.42 0.43 – 0.45 0.18 – 0.26 1.32 – 1.38 1.15 – 1.21 1.15 – 1.29 0.54 – 0.85 0.70

0.047

Note: Values were converted from Euro to US$ (1 Euro = 1.20 US$ as of 27 April 2007)

0.195

0.74

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COMMENTS Arid and semi-arid regions are facing increasingly more serious water shortage problems. As the population grows in these areas, water is an increasingly valuable and limited resource. Every effort must be made to use water more efficiently, and new practices are being developed and implemented in the field of water use and water conservation [27,28]. Hemodialysis represents an environmental challenge, in part due to high water consumption [8]. In regions with water scarcity, high consumption of water during by hemodialysis units is a compelling argument supporting wastewater recycling. This paper discusses the technical and economic feasibility of recycling this type of wastewater for potential use in irrigation. Observed values show that organic matters and bacterial biomass were under the acceptable limits; however, conductivity values exceeded FAO standards [15]. Due to this high conductivity, wastewater must be treated to accepted standards prior to use for irrigation [22,23]. Membrane separation has proven to be the preferred treatment process for such high conductivity wastewaters [22,23,29]. Moreover, this technology has previously been shown to be efficient and economical in comparison with other approaches [24,25,30,31]. In this study, computer simulations were executed with two models of membrane treatment (nanofiltration and reverse osmosis) based on the characteristics of the influent and the desired effluent wastewater quality and assumptions obtained from prior literature [30,32]. The simulations suggested that both methods showed greater benefit compared to desalination of seawater, resulting in a cost savings (or benefit) of 20 - 30%. Membrane separation is a widely used process for the treatment of various types of wastewater. However, to our knowledge, application of this technology to hemodialysis wastewater has not been performed. Also, in reviewing the literature, we were unable to document any engineering application system related to hemodialysis wastewater treatment. Consequently the result of our analysis calls for further investigations in this area. In conclusion, due to the high water consumption in hemodialysis, it is essential to study its potential for recycling. Through analysis and evaluation of the technical and economic feasibility of hemodialysis wastewater treatment, this study draws attention to this important but neglected aspect of hemodialysis therapy.

REFERENCES [1] [2] [3] [4]

Rosegrant MW, Cai X, Cline SA. World Water and Food to 2025: Dealing With Scarcity. Washington DC, International Food Policy Research Institute, 2002 Berrittella M, Hoekstra AY, Rehdanz K, Roson R, Tol RS. The economic impact of restricted water supply: a computable general equilibrium analysis. Water Res. 2007; 41: 1799-813. Moe CL, Rheingans RD. Global challenges in water, sanitation and health. J Water Health. 2006; 4 Suppl 1: S41-S57. Tal A. Seeking sustainability: Israel's evolving water management strategy. Science. 2006; 313: 1081-4.

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Faissal Tarrass, Meryem Benjelloun, and Omar Benjelloun Gaan N. Environmental scarcity of water and Indo-Nepal conflict: towards environmental integration and cooperation. Asian Profile 2001; 29: 417-28. Wang XC, Jin PK. Water shortage and needs for wastewater re-use in the north China. Water Sci Technol 2006; 53: 35-44. Lusamvuku A, Hermelin-Jobet I, Boudard B, Bracquemont MC, Lebas J. Hemodialysis water production: evaluation and assurance quality management. J Pharm Clin 1999; 18: 300-5. Cousin P. Traitement d'eau en hémodialyse. Stage de Perfectionnement à l'Ingénierie Hospitalière, Université de Technologie de Compiègne 1998-99, pp 59. http://www.utc.fr/~farges/spibh/98-99/Stages/Cousin/Cousin.htm, Assessed : 28th February 2008. Moroccan Ministry of Health. Strategy of care from 2008 to 2012. Available at: http://www.sante.gov.ma/Leministre/Communique/communique2008/corruption.htm, Accessed: 5th March 2008. Kingdom of Morocco Water Sector Review. Report No 14750-MOR. Washington DC, World Bank, 1995. Pauwels B, Verstraete W. The treatment of hospital wastewater: an appraisal. J Water Health. 2006; 4: 405-16. American Public Health Association. Standard methods for the examination of water and wastewater. 20th ed. Washington: Washington DC, American Water Works Association, Water Environment Federation, 1998. Association for the Advancement of Medical Instrumentation. American National Standard Hemodialysis Systems. ANSI/AAMI RD5-1993. Arlington, VA: AAMI; 1993. Carr RM, Blumenthal UJ, Mara DD. Guidelines for the safe use of wastewater in agriculture: revisiting WHO guidelines. Water Sci Technol 2004; 50: 31-8. Food and Agriculture Organization of the United Nations. Wastewater treatment and use in agriculture, Rome, FAO, 1992. Moch I. Development of a CD-ROM cost program for water treatment projects, Memb Technol 2003; 6: 5-8. Food and Agriculture Organization of the United Nations. Water desalination for agricultural applications, Rome, FAO, 2006. Asano T, Smith RG, Tchobanoglous G: Municipal wastewater: Treatment and reclaimed water characteristics. in Pettygrove GS, Asano T (eds): Irrigation With Reclaimed Municipal Wastewater—A Guidance Manual. Chelsea, MS, Lewis, 1985, pp 1-26. Soudi B, Xanthoulis D. Projet de gestion des ressources en eau: Élaboration des dossiers techniques relatifs aux valeurs limites des rejets industriels dans le domaine public hydraulique. Rome, FAO, 2006. Cotruvo JA. Water Desalination Processes and Associated Health and Environmental Issues. Water Cond Purif 2005; 47: 13-7. Gupta AK, Gupta SK, Patil RS. Statistical analyses of coastal water quality for a port and harbour region in India. Environ Monit Assess 2005; 102: 179-200. Magesana GN, Williamsona JC, Yeatesb GW, Lloyd-Jones AR. Wastewater C:N ratio effects on soil hydraulic conductivity and potential mechanisms for recovery. Bioresource Technol 2000; 71: 21-27.

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[23] Gerhart VJ, Kane R, Glenn EP. Recycling industrial saline wastewater for landscape irrigation in a desert urban area. J Arid Environ 2006; 67: 473-86. [24] Moch I, Chapman M, Steward D. Estimating membrane water treatment costs Membr Technol 2003; 8: 5-7. [25] Cote P, Masini M, Mourato D. Comparison of membrane options for water reuse and reclamation. Desalination 2004; 167: 1-11. [26] Office National de l‘Electricité: Customer space. Available at: www.one.org.ma. Accessed 28th February 2008. [27] Bakir HA. Sustainable wastewater management for small communities in the Middle East and North Africa. J Envir Manag 2001; 61: 319-28. [28] Abu-Zeid KM. Recent trends and developments: reuse of wastewater in agriculture. Envir Manag Health 1998; 2: 79-89. [29] Marcucci M, Ciabatti I, Matteucci A, Vernaglione G. Membrane technologies applied to textile wastewater treatment. Ann NY Acad Sci 2003; 984: 53-64. [30] Noronha M, Mavrov V, Chmiel H. Simulation model for optimisation of two-stage membrane filtration plants; minimising the specific costs of power consumption. J Membr Sci 2002; 202: 217-32. [31] Hafez A, Khedr M, Gadallah H. Wastewater treatment and water reuse of food processing industries. Part II: Techno-economic study of a membrane separation technique. Desalination 2007; 214: 261-72. [32] Moch I, Chapman M, Steward D. Development of a CD-ROM cost program for water treatment projects. Membr Technol 2003; 6: 5-8.

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ISBN: 978-1-60741-915-0 © 2010 Nova Science Publishers, Inc.

Chapter 16

PB (II) IONS REMOVAL BY DRIED RHIZOPUS OLIGOSPORUS BIOMASS PRODUCED FROM FOOD PROCESSING WASTEWATER 1

H. Duygu Ozsoy1* and J. Hans van Leeuwen2,3,4 Dept. of Environmental Engineering, Engineering Faculty, Mersin University,

33340, Mersin, Turkey. Dept. of Civil, Construction and Environmental Engineering. 3 Dept. of Agricultural and Biosystems Engineering. 4 Dept. of Food Science and Human Nutrition, Iowa State University, Ames, IA 50011, USA. 2

ABSTRACT Heavy metal pollution is a serious problem in many developed and developing countries. Lead had been recognized as a particularly toxic metal and comes into water bodies mainly from metallurgical, battery, metal plating, mining and alloy industries. In order to minimize the impacts of this metal on human health, animals and the environment, lead-contaminated water and wastewater need to be treated before discharge to water bodies. This chapter concerns an investigation of potential usage of corn-processing wastewater as a new alternative low-cost substrate to produce biosorbent and evaluate this biosorbent to remove Pb(II) ions from aqueous solutions. For this aim, Rhizopus oligosporus cultivated on corn-processing wastewater and dried biomass of these fungi was used as an adsorbent. The adsorption experiments were conducted in a batch process and the effects of contact time (1-48 hours), initial pH (2-7), initial metal ion concentration (20-100 mg L-1) and adsorbent dosage (0.5-5 g L-1) on the adsorption were investigated. Pb (II) ion concentrations before and after adsorption were measured using Inductively Coupled Plasma-Mass Spectrometry. Maximum adsorption capacity was achieved at pH 5.0. The isothermal data of dried fungal biomass could be described well by the Langmuir equation and monolayer capacity had a mean value of 59.88 mg g-1. The *

Corresponding author: Phone: +90 324 361 00 01/7102, Fax: +90 324 361 00 99, E-mail: [email protected] (H.D. Ozsoy)

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H. Duygu Ozsoy and J. Hans van Leeuwen pseudo-second order reaction model provided the best description of the data with a correlation coefficient 0.99 for different initial metal concentrations. This result indicates that chemical sorption might be the basic mechanism for this adsorption process and Fourier Transform Infrared Spectroscopy analyses showed that amide I and hydroxyl groups play an important role in binding Pb (II). Because of the high activation capacity of adsorbent and low cost of process dried R. oligosporus biomass presents a good potential as an alternative material for removal of Pb (II) ions from the aqueous solutions.

Keywords: Biosorption, Food processing wastewater, Heavy metals, Pb (II), Rhizopus oligosporus.

1. INTRODUCTION Removal of heavy metals from aqueous solutions is one of the major problems in industrial wastewater treatment because most of them are toxic even at very low concentrations. The amount of these pollutants in water has been increased with industrial applications including mining, refining, electroplating and production of textiles, paints and dyes [1]. Lead had been recognized as a particularly toxic hazardous environmental pollutant and comes into water bodies mainly from metallurgical, battery, metal plating, lead smelting, mining and alloy industries. In order to minimize the impacts of this metal on human health, animals and the environment, lead-contaminated water and wastewater need to be treated before discharge to water bodies [2]. The conventional methods for removing metals from aqueous solutions include chemical precipitation, chemical oxidation or reduction, electrochemical treatment, reverse osmosis, solvent extraction, ion exchange and evaporation. However, these techniques have several disadvantages such as high chemical cost, low removal efficiency, low selectivity, high energy requirements, and generation of secondary toxic slurries [3-5]. Therefore removal of toxic heavy metals in a cost-effective and environment-friendly manner assumes great importance. Adsorption is a highly effective and economical technique for removal of heavy metals from aqueous solutions. Commercial activated carbon is a well-known and highly effective adsorbent, but the high cost of activated carbon limits its use as an adsorbent especially in developing countries [6]. From this standpoint, numerous investigations were conducted by scientists in this growing and important field of research for the exploration of alternative methods using less expensive natural materials [7]. Metal-sorption by various types of biomaterials like metabolically inactive dried biomass of algae, bacteria and fungi can find useful application for removing metals from solution because of their unique chemical composition [8-11]. Research indicated that biosorption is a very effective method to remove metals from the water and wastewater. Cultivation of microorganisms requires a bioreactor and nutrients such as carbon, nitrogen and trace elements [12,13]. Therefore, cultivation cost is the most important factor to produce these biosorbents. This chapter presents experimental results on removal of Pb (II) ions from aqueous solution by dried fungal biomass produced from corn-processing wastewater. These results

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show that it is possible to use food-processing wastewater as a substrate for cultivating the fungal biomass to reduce operational costs of adsorption processes. Using the wastewaters would be particularly attractive and cost effective because there are many food-processing plants in USA and many other countries that could provide suitable industrial wastewater for cultivating the microbial biomass, such as fungi. The wastewater needs to be treated to address discharge objectives and therefore, the biomass is produces at minimal cost. The goal of the method was to investigate the efficacy of dried fungal biomass as adsorbent for removal of Pb (II) ions from aqueous solution and reduce the treatment costs using another wastewater as a substrate for fungi.

2. EXPERIMENTAL The experimental program comprised two phases. (a) Rhizopus oligosporus were grown on food-processing wastewater. (b) The harvested and dried fungi used for adsorption of Pb (II) ions.

2.1. Cultivation of Rhizopus Oligosporus Rhizopus oligosporus was obtained from American Type Culture Collection (Rockville, MD). The culture was rehydrated and revived in yeast-malt (YM) nutrient broth at 24 ºC. The revived culture was transferred on to numerous potato dextrose agar (PDA) plates and incubated at room temperature (24 ºC) for 7 days. Then fungal sporangiospores were harvested from the surface of PDA plates into sterile distilled water containing 0.85 % (w/v) saline solution (NaCl) and 0.5 % (v/v) of Tween 80. The harvested cultures were diluted further to achieve a spore count of 106 to 107 spores/mL, determined by haemocytometer counts. Glycerin (20 %; v/v) was added to the spore suspension as a cryoprotectant for ultralow frozen storage at –75 oC in 2 mL cryo-vials for future use as a bioreactor inoculum. The inocula were used as a seed in laboratory-scale continuous attached growth tank reactors using corn-processing wastewater as organic substrate. The wastewater was supplied from the ADM wet corn milling facility in Cedar Rapids, IA, US. The reactors were operated at a hydraulic retention time (HRT) of 8 h and solids retention time (SRT) of 2 days. These HRT and SRT values were found to be optimal for the maximum growth of the Rhizopus oligosporus [14]. The micro-fungi were growing in the form of attached mycelia and harvested daily from the bioreactor by natural sloughing off the attachment surface and subsequent gravity settling. The mycelia were washed with deionised water and dried at 65 ºC for 24 h. The dried fungal pellets were ground and sieved (0.5 mm< diameter).

2.2. Adsorption of Heavy Metals The effects of contact time, initial pH, initial metal ion concentration and temperature on adsorption efficiency were examined through a series of shaker flask tests. After determining

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the optimum conditions, a series of adsorption tests were conducted to determine the isotherm for Pb (II) ions.

2.3. Chemicals A 1 g L-1 stock solution of lead was prepared with single reagent grade metal solution (Claritas, Fisher Chemicals) in deionized water. The metal solution was diluted to appropriate concentrations as needed and stored at 4 oC until further use. HNO3 and NaOH were obtained from Fisher Chemicals and used for pH value adjustment.

2.4. Adsorption Experiments All sorption tests were conducted using single reagent grade metal to minimize the variability of metal concentrations and to avoid competitive adsorption of mixed metals on adsorbent. 100 mL of metal solution was added to each of flask containing 0.1 g (dry weight) of R. oligosporus. The flasks were placed on an orbital shaker table running at 150 rpm at 30±1 C (except the temperature experiments) until equilibrium was reached. The residual concentration of Pb (II) ions in the aqueous phase (obtained by centrifugation, 1000 g-10 min) was determined using Inductively Coupled Plasma-Mass Spectrometry (ICP-MS). All tests were conducted in triplicate. The concentrations of the Pb (II) ions the in aqueous phase were used to determine the adsorption capacity of R. oligosporus. Equilibrium sorption isotherms were determined by mass balance. The amount of adsorbed Pb (II) ions at equilibrium, qeq (mg g-1) was calculated as follows: qeq =

[(Co Ceq )V ] x

(1)

where Co and Ceq are the initial and equilibrium concentrations of Pb (II) ions (mg L-1), V volume of solution and x the weight of sorbent (g).

2.5. Inductively Coupled Plasma-Mass Spectrometry (ICP-MS) Analysis Measurements were performed with a Hewlett Packard 4500 Series ICP-MS using external calibration. The instrument was calibrated before each measurement. Operating parameters are summarized in Table 1.

2.6. Equilibrium Isotherms and Kinetics of Adsorption The Langmuir isotherm was used first to describe observed sorption phenomena. The Langmuir isotherm applies to adsorption on a completely homogenous surface with negligible interaction between adsorbed molecules [15,16]. The linear form of the equation can be written as;

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Table 1. ICP-MS operating conditions. Rf power Rf matching Sample depth Plasma gas flow Auxiliary gas flow

1200 W 2V 7.8 mm 16 L min-1 1.4 L min-

Carrier gas flow

1.0 L min-

Acquisition time Resolution

22.83 sec 300

Ceq qeq

=

1 1

Ceq 1 + bqmax qmax

(2)

where Ceq is the equilibrium concentration of Pb (II) ions, qeq is the amount of adsorption at equilibrium, qmax is the mono-layer capacity, and b is an equilibrium constant of Langmuir. The Freundlich isotherm (empirical model adsorption in aqueous systems) was also tested with our experimental data. The linear form of the equation can be written as: lnqeq= lnKf +

1 n

lnCeq

(3)

where Kf is the measure of sorption capacity, 1/n is the sorption intensity. Pseudo-first order and pseudo-second order kinetic models were applied to data to analyse the sorption kinetics of Pb (II) ions. A simple pseudo first-order equation due to Lagergren was used by different researchers [17,18]: log (qeq-qt) = log qeq-

kad 2.303t

(4)

where qe and qt are the amount of adsorption at equilibrium and at time t respectively, and kad is the rate constant of the pseudo first-order adsorption process. A plot of log (qeq-qt) vs. t would provide a straight line for first-order adsorption kinetics, allowing computation of the adsorption rate constant, kad. Ho‘s second-order rate equation, which has been called a pseudo-second order kinetic expression, has also been applied widely [19,20]. The linear form of the kinetic rate equations can be written as follows:

t qt

1 kqeq2

1 t qe q

(5)

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where k is the rate constant of sorption (dm3 mg-1 min-1), qe is the amount of metal ion sorbed at equilibrium (mg g-1), and qt is the amount of metal ion sorbed at time t (mgg-1). The constants can be determined experimentally by plotting t/qt against t.

2.7. FT-IR Analyses Fourier Transform Infrared Spectroscopy (FT-IR) analysis in the solid phase was performed using a Fourier Transform Infrared Spectrometer (Varian 2000 FT-IR). Pure biosorbent powders were used and spectra of the fungal biomass before and after Pb(II) sorption were recorded.

3. RESULTS 3.1. The Effect of the Contact Time The major fraction of Pb (II) ions, adsorbed within the first 6h and dissolve Pb remained nearly constant afterwards. This suggested that the biosorption process reached saturation within 6h. For this reason a 6h contact time was used for the further experiments. Figure 1 shows the effect of contact time on adsorption of Pb(II) ions onto the dried R. oligosporus biomass.

3.2. The Effect of the Initial pH Results of the experiments using 100 mg L-1 Pb (II) solutions and 1 g L-1 adsorbent showed that efficiencies of adsorption were increased with increasing pH from 2.0 to 6.0 (Figure 2). At the low pH ranges, a high concentration of protons in the solution may have competed with metal ions in forming a bond with the active sites on the surface of the fungi. These bonded active sites thereafter became saturated and were unavailable to other cations. 60.00

q (mg/g)

50.00 40.00 30.00 20.00 10.00 0.00 0

5

10

15

20

25

30

35

40

45

50

55

t (hour)

Figure 1. The effects of contact time on adsorption of Pb(II) ions (100 mg l-1) to the dried R.oligosporus biomass (adsorbent dosage: 1g l-1; pH:5.0; temperature: 30 C).

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60.00

q (mg/g)

50.00 pH2 pH3 pH4 pH5 pH6

40.00 30.00 20.00 10.00 0.00 0

1

2

3

4

5

6

7

t (hour)

Figure 2. The effects of the initial pH on adsorption of Pb(II) ions (100 mg l-1) to the dried R.oligosporus biomass (adsorbent dosage:1g l-1; contact time: 6h; temperature: 30 C) 60

q(mg/g)

50 40 30 20 10 0 0

20

40

60

80

100

120

140

160

Co(mg/L)

Figure 3. The effects of initial metal concentration on adsorption of Pb(II) ions to the dead R.oligosporus biomass (adsorbent dosage:1g l-1; contact time: 6h; pH:5.0).

3.3. The Effect of the Initial Metal Concentrations The adsorption of Pb (II) by the dried R. oligosporus biomass was studied at different Pb (II) ion concentrations in the range from 20 to 150 mg L-1. Equilibrium sorption capacity of the dried R. oligosporus biomass increased with increasing initial Pb (II) ion concentrations (Figure 3). The initial concentration provides an important driving force to overcome all mass transfer resistance of Pb (II) ions between the aqueous and solid phases. Hence a higher initial concentration of Pb (II) ions may increase the adsorption capacity.

3.4. The Effect of Adsorbent Dosage Experimental results indicated that the efficiency of biosorption was decreased with increasing adsorbent dosage ranging from 0.5-5 g L-1 (Figure 4). This However the percentage of the Pb(II) ions biosorption was increased with increasing biomass because of higher surface area. This can be explained by concentration gradient between the solute

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concentration in the solution and the one in the surface of the adsorbent. When the adsorbent dosage is higher, there is a very fast adsorption onto the adsorbent surface, which results in a lower Pb(II) ion concentration in the solution. However, the adsorption sites on the adsorbent surface remain unsaturated when the Pb(II) ion concentration in the solution drops to a lower value. Thus, the amount of Pb(II) ions adsorbed onto per unit weight of adsorbent gets reduced with the adsorbent dosage increasing [25].

3.5. Adsorption Isotherms and Kinetics of Adsorption

q (mg/g)

Two equilibrium models were employed: The Langmuir and Freundlich isotherm equations. The correlation coefficient of Freundlich isotherm (R2) was 0.8824 (Figure 5). The Langmuir model was the best-fit isotherm for adsorption of Pb (II) to the dried R. oligosporus biomass. Langmuir isotherm model parameters, qmax and b, were estimated from the intercept and slope of Ceq/qeq vs. Ceq , according to Eq. (2) and obtained as 59.88 (mg g-1) and 0.042 (L g-1), respectively. The correlation coefficient of the Langmuir isotherm (R2) was 0.9820 (Figure 6). 90 80 70 60 50 40 30 20 10 0

0.5 1 3 5

0

1

2

3 4 t (hour)

5

6

7

lnqe

Figure 4. The effects of adsorbent dosage on adsorption of Pb (II) ions to dried R. oligosporus biomass (contact time: 6h; pH:5.0; temperature: 30 C) 4.5 4 3.5 3 2.5 2 1.5 1 0.5 0

y = 0.4115x + 2.0629 R2 = 0.8824

2

2.5

3

3.5

4

4.5

5

lnCe

Figure 5. The Freundlich adsorption isotherm for Pb(II) adsorption on to dried R. oligosporus biomass (adsorbent dosage:1g l-1; pH: 5.0; initial metal concentration of 20-100 mg l-1; temperature: 30 C).

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2.500 y = 0.0167x + 0.398 R2 = 0.982

Ce/qe

2.000 1.500 1.000 0.500 0.000 0

20

40

60

80

100

120

Ce (mg/L)

Figure 6. The Langmuir adsorption isotherm for Pb(II) adsorption on to dried R. oligosporus biomass (adsorbent dosage:1g l-1; pH 5.0; initial metal concentration of 20-100 mg l-1; temperature: 30 C). 20 mg/L 40 mg/L 60 mg/L 80 mg/L 100 mg/L

25.00

t/qt

20.00 15.00

R2 = 0.995 R2 = 0.9982 R2 = 0.9929

10.00

R2 = 0.9924 R2 = 0.9981

5.00 0.00 0

100

200 t (min)

300

400

Figure 7. The plots of pseudo-second order kinetics with respect to different initial Pb(II) ion concentrations (adsorbent dosage:1g l-1; pH 5.0; temperature:30 C).

Kinetic studies were carried out for biosorption of Pb (II) as a function of contact time at various initial Pb (II) concentrations ranging from 20-100 mg L-1. Experimental results indicated that the pseudo-second order reaction model provided the best description of the data with a correlation coefficient 0.99 for different initial metal concentrations (Figure 7). Table 2. Pseudo-second order reaction rate regression results. Initial metal concentration (mg L-1) 20 40 60 80 100

Rate constant (L mg-1 min1 ) 8.9 x10-4 5.0 x10-4 3.9 x10-4 3.8 x10-4 3.7 x10-4

Correlation coefficient (R2) 0.9950 0.9982 0.9929 0.9924 0.9981

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Figure 8. FTIR spectra of the dried R. oligosporus biomass.

Figure 9. FTIR spectra of the dried R. oligosporus biomass after Pb(II) ions adsorption.

Reaction rate constants for pseudo-second order equations are shown at Table 2. The results indicated that the adsorption system studied follows a pseudo-second order kinetic model at all time intervals and pseudo-second order rate constants were affected by initial Pb (II) ions concentration.

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3.6. FT-IR Analysis Fungal biomass is a complex material containing protein, lipid and polysaccharides (cellulose, chitin) as major constituents. Therefore biosorption may occur at the polar functional groups of cellulose and chitin, which include amino, hydroxyl and carboxyl groups as chemical bonding agents. The FTIR analyses of dried R. oligosporus biomass before and after Pb(II) adsorption is shown in Figure 8-9. Spectra analyses of FTIR spectrum showed that there was a decrease in the adsorption intensity of amide I and –OH groups at 1644 cm-1, 1402cm-1 respectively and this indicates that amide I and hydroxyl groups played an important role in binding Pb(II).

4. CONCLUSION To development of an efficient and cost-effective removal process, fungal biomass produced from food industry wastewater is a good alternative biosorbent. Experimental results showed that based on the Langmuir coefficients, the total capacity (monolayer saturation at equilibrium) of the dried R. oligosporus biomass for Pb (II) ions was about 60 mg g-1 (biosorbent dose of 1g/L, 6h contact time, initial Pb (II) concentration of 100 mg/L and optimum pH of 5.0). Experimental results also indicated that the pseudo-second order reaction model provided the best description of the data with a correlation coefficient 0.99 for different initial metal concentrations. The fit of the experimental data to this model suggest that the process controlling the rate may be chemical sorption. The FTIR analyses showed that amide I and hydroxyl groups plays an important role in binding of Pb(II). With the advantage of high metal biosorption capacity, R.oligosporus biomass, produced from corn-processing wastewater, has the potential to be used as an effective and economic biosorbent material for the removal of Pb (II) ions from wastewater streams.

ACKNOWLEDGMENTS Authors are thankful to Dr. Basudeb Saha for technical assistance (ICP-MS measurements) and Iowa State University for their financial support.

REFERENCES [1] [2] [3]

Aguado, J., Arsuaga, J. M., Arencibia, A., Lindo, M. & Gascon, V. (2009). Aqueous heavy metals removal by adsorption on amine-functionalized mesoporous silica. Journal of Hazardous Materials, 163, 213-221. Yurtsever, M. & Sengil, I. A. (2009). Biosorption of Pb(II) ions by modified quebracho tannin resin, Journal of Hazardous Materials, 163, 58-64. Ozsoy, H. D. & Kumbur, H. (2006). Adsorption of Cu(II) ions on cotton boll. Journal of Hazardous Materials, B136, 911-916.

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H. Duygu Ozsoy and J. Hans van Leeuwen Rakhshaee, R, Giahi, M. & Pourahmad, A. (2009). Studying effect of cell wall‘s carboxyl-carboxylate ratio change of Lemna minor to remove heavy metals from aqueous solution. Journal of Hazardous Materials, 163, 165-173. Gupta, V. K. & Rastogi, A. (2009). Biosorption of hexavalent chromium by raw and acid-treated green alga Oedogonium hatei from aqueous solution. Journal of Hazardous Materials, 163, 396-402. Boujelben, N., Bouzid, J. & Elouear, Z. (2009). Adsorption of nickel and copper onto natural iron oxide-coated sand from aqueous solutions: Study in single and binary systems, Journal of Hazardous Materials, 163, 376-382. Aziz, A., Ouali, M. S., Elandaloussi, El. H., De Menorval, L. C. & Lindheimer, M. (2009). Chemically modified olive stone: a low-cost sorbent for heavy metals and basic dyes removal from aqueous solutions. Journal of Hazardous Materials, 163, 441-447. Ahluwalia, S. S. & Goyal, D. (2007). Microbial and plant derived biomass for removal of heavy metals from wastewater. Bioresource Technology, 98(12), 2243-2257. Ozsoy, H. D, Kumbur, H & Ozer, Z. (2007). Adsorption of copper (II) ions to peanut hulls and Pinus brutia sawdust. International Journal of Environmental Pollution, 31, 125-134. 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. Ozsoy, H. D, Kumbur, H., Saha, B. & van Leeuwen, J. H. (2008). Use of Rhizopus oligosporus produced from food processing wastewater as a biosorbent for Cu(II) ions removal from the aqueous solutions. Bioresource Technology, 99, 4943-4948. Yurekli, F., Unyayar, A., Porgali, Z. B. & Mazmanci, M. A. (2004). Effects of cadmium exposure on phytochelatin and the synthesis of abscisic acid in Funalia trogii. Engineering in Life Sciences, 4, 378-380. Mazmanci, M. A. & Unyayar, A. (2005) Decolorisation of Reactive Black 5 by Funalia trogii immobilized on Luffa cylindirica sponge. Process Biochemistry, 40, 337-342. Jasti, N., Khanal, S. K., Pometto III, A. L. & van Leeuwen, J. H. (2008). Converting corn wet milling effluent into high-value fungal biomass in an attached growth bioreactor. Biotechnology & Bioengineering, 101(6), 1223-1233. Yu, B., Zhang, Y., Shukla, A., Shukla, S. S. & Dorris, K. L. (2001). The removal of heavy metals from aqueous solutions by sawdust adsorption-removal of lead and comparison of its adsorption with copper. Journal of Hazardous Materials, B84, 83-94. Ho, Y. S. (2004). Pseudo-isotherms using a second order kinetic expression constant. Adsorption, 10, 151-158. Ho, Y. S. & McKay, G. (1998). Sorption of dye from aqueous solution by peat. Chemical Engineering Journal, 70, 115-124. Bhattacharyya, K. G. & Sharma, A. (2005). Kinetics and thermodynamics of methylene blue adsorption on neem (Azadirachta Indica) leaf powder. Dyes and Pigments, 65, 51-59. Ho Y. S. & McKay, G. (2000). The kinetics of sorption of divalent metal ions onto sphagnum moss peat. Water Resource, 34(3), 735-742. Ho, Y. S. (2005). Effect of pH on lead removal from water using tree fern as the sorbent. Bioresource Technology, 96, 1292-1296.

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[21] Fiol, N., Villaescusa, I., Martinez, M., Miralles, N., Poch, J. & Serarols, J. (2006). Sorption of Pb(II), Ni(II), Cu(II) and Cd(II) from aqueous solutions by olive stone waste. Separation and Purification Technology, 50, 132-140. [22] Pavasant, P., Apiratikul, R., Sungkhum, V., Suthiparinyanont, P., Wattanachira, S. & Marhaba, T. F. (2006). Biosorpion of Cu2+, Cd2+, Pb2+ and Zn2+ usind dried marine green macroalga Caulerpa lentillifera. Bioresource Technology, 97, 2321-2329. [23] Liu, Y., Chang, X., Guo, Y. & Meng, S. (2006). Biosorption and preconcentration of lead and cadmium on waste Chinese herb Pang Da Hai. Journal of Hazardous Materials, B135, 389-394. [24] Martinez, M., Miralles, N., Hidalgo, S., Fiol, N., Villaescusa, I. & Poch, N. (2006). Removal of lead(II) and cadmium(II) from aqueous solutions using grape stalk waste. Journal of Hazardous Materials, B133, 203-211. [25] Jiang, Y., Pang, H. & Liao B. (2009) Removal of copper(II) ions from aqueous solution by modified bagasse. Journal of Hazardous Materials, 164, 1-9.

In: Fluid Waste Disposal Editor: Kay W. Canton, pp. 333-353

ISBN: 978-1-60741-915-0 © 2010 Nova Science Publishers, Inc.

Chapter 17

CONTROL OF PLASTICIZERS IN DRINKING WATER, EFFLUENTS AND SURFACE WATERS Rosa Mosteo, Judith Sarasa, Mª Peña Ormad and Jose Luis Ovelleiro Department of Chemical Engineering and Environmental Technologies, University of Zaragoza-Spain.

ABSTRACT The main objective of this research work is to determine the presence of di(2ethylhexyl) phthalate, di(2-ethylhexyl) adipate and diisodecyl phthalate, in different water samples (drinking waters, effluents and surface waters). Different analytical methods were studied in order to know the best methodology for the quantification of these compounds. Solid-liquid and liquid-liquid extraction were investigated and finally the liquid-liquid extraction and analysis by gas chromatography followed by mass spectroscopy was chosen because of offering the highest recovery rate. In the whole of this research study, the control of background pollution by reagents and material was extremely important. The problem of background pollution is more serious in the trace analysis of phthalates and adipates as a consequence of their presence in almost all equipment and reagents used in the laboratory. Respect to the control of the selected plasticizers in the different water samples, bis (2-ethylhexyl) phthalate and bis (2-ethylhexyl) adipate were detected in drinking water, effluents and surface waters. On the other hand, diisodecyl phthalate was not detected in any sample.

INTRODUCTION Phthalates have been in use for almost 40 years and are used in the manufacture of PVC and other resins, as well as plasticizers and insect repellents [1]. Plasticizers are used in building materials, home furnishing, transportation, clothing, and, to a limited extent, in food packaging and medicinal products [2]. There is also concern regarding the potential health

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effects of several phthalates because these compounds are used to impart softness and flexibility to normally rigid PVC products in children‘s toys. Phthalates can enter the environment through losses during manufacture and by leaching from the final product. This is because they are not chemically bonded to the polymeric matrix [1]. These compounds have low water solubilities and tend to adsorb to sediments and suspended solids. Therefore, they have the potential to leach into their surrounding environment. Certain phthalates have also shown estrogenic behaviour [3, 4] so they can be classified as potential endocrine disruptors. Many endocrine disrupting substances, or potential endocrine disrupters (EDCs), were previously classified as organic micropollutants. Therefore, many substances, such as foodstuffs, flavonoids, lignans, sterols, fungal metabolites and synthetic chemicals of widely varying structural classes (e.g. phthalates, PCBs), can interact with hormone receptors and modulate the endocrine system [5]. Phthalates are classified as EDCs in several lists of compounds by various organizations [6], for example: -

UKEA: United Kingdom Economics Account [7]. USEPA: U.S. Environmental Protection Agency [8]. OSPAR: Oslo and Paris Commissions [9].

The most of Phthalates and adipates used as plasticizers are being included in list of priority contaminants of different countries. In the United Stated, the Environmental Protection Agency established concentration limits in drinking waters for both di(2ethylkhexyl)phthalate (6 µg/l) and di(2-ethylhexyl)adipate (0.4 mg/l) [10]. In the European Union there are not contamination limits for phthalates and adipates, although it is known that these limits will be established in the near future. The di(2-ethylkhexyl)phthalate is included in the priority list [11] and identified as priority hazardous substance in the field of water policy. The guideline value in drinking water proposed by the World Health Organization in the 1993 Guidelines [12, 14] for the Di(2-ethylhexyl)adipate is 80 µg/l and in the case of Di(2-ethylhexyl)phthalate the guideline value is 8 µg/l. Di(2-ethylhexyl)adipate is also known as DEHA. This compound is mainly used as a plasticizer for synthetic resins such as PVC, but significant amounts are also used as lubricants and for hydraulic fluids [13]. Reports of the presence of DEHA in surface water and drinking water are scarce, but DEHA has occasionally been identified in drinking-water at levels of a few micrograms per litre. As a consequence of its use in PVC films, food is the most important source of human exposure (up to 20 mg/day). Di(2-ethylhexyl)phthalate is also known as DEHP. This compound is primarily used as a plasticizer in many flexible polyvinyl chloride products and in vinyl chloride co-polymer resins. It has also application as replacement for polychlorinated biphenyls in dielectric fluids for small (low-voltage) electrical capacitors [13]. This compound has been found in surface water, groundwater and drinking-water in concentrations of a few micrograms per litre. In polluted surface water and groundwater, concentrations of hundreds of micrograms per litre have been reported. Numerous manufactures are trying to substitute the DEHP for the diisodecyl phthalate, which is a plasticizer less toxic than di(2-ethylhexyl) phthalate (DIDP). As a consequence of the necessary control of these pollutants in different waters, analytical methodologies should be established in order to improve the quantification of these compounds. Methodologies commonly used to analyze organic compounds in trace levels are based on liquid-liquid or solid-liquid extraction following chromatographic analysis. In this

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study, two methodologies have been developed in order to find the best one for the analysis of the interest compounds. Both of them are modifications of the 506 and 606 EPA methods [15, 16]. In this research work the presence of bis (2-ethylhexyl) phthalate, bis (2-ethylhexyl) adipate and diisodecyl phthalate, plasticizers frequently used by manufacturers, has been evaluated in different water samples (drinking waters, effluents and surface waters). Different analytical methodologies, using solid-liquid extraction or liquid-liquid extraction prior to analysis by GC/MS, were studied in order to know the best methodology for the quantification of these compounds. A study about the background pollution is carried out since they are present in the majority of equipment and reagents used in the laboratory.

EXPERIMENTAL PROCEDURE Analytical Methodology for the Control of DEHP, DIDP and DEHA Standard of DEHP, DIDP and DEHA purchased from Dr. Ehrenstorfer were used. A stock solution for each compound is prepared in methanol. In the case of the DEHP, the stock solution presented concentration of 1000 mg/l, the DIDP 1520 mg/l and the DEHA 1000 mg/l. The chemical structure of DEHP, DIDP and DEHA are shown in Figure 1. The chromatographic conditions are shown in Table 1. DEH P

DEH A

DIDP

Figure 1.Chemical structure of DEHP, DEHA and DIDP.

Table 1. Chromatographic conditions. Gas Chromatographer TRACE GC 2000 (TermoFinnigan) Column DB5 (J&W, 30 m, 0,25 mm, 0,25 μm) Program of temperature 60 ºC (1.5 min)-10 ºC min-1-240 ºC (30 min)-10 ºC min-1-260 ºC (5 min) Temperature of injector 300 ºC Volume of injection 2 μL, splitless 0.8 min Carrier gas He, 1 mL min-1 Mass Espectrometer POLARIS (ThermoFinnigan) Energy of ionization 70 eV Mode of acquisition SIM Range of masses 50-450 amu Velocity of screened 1 scan s-1

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Rosa Mosteo, Judith Sarasa, Mª Peña Ormad et al. 506 and 606 U.S EPA methods

506 U.S EPA method Sample (4ºC)

Sample (4ºC)

Sample dilution with MIlliQ water (if necessary) 506 Method + Na Cl

Liquid-liquid extraction 3 extactions with CH2Cl2 506 method:+1extaction C6H6

Concentration with N2 in TurboVap

Sample dilution with MIlliQ water (if necessary)

Solid-liquid extraction C18 cartridge

- Cartridge preparation (CH2Cl2/MeOH/H2O) - Compound adsorption - Cartridge dried with air

Elution of compounds with CH2Cl2 (5 ml, pressure)

Extract dried with Na2SO4 anhidrous

Extract dried with Na2SO4 anhidrous

Concentration with N2 (volume 1 ml)

Concentration with N2 (volume 1 ml)

GC/MS analysis

GC/MS analysis

Figure 2. Diagrams of liquid-liquid extraction (506 and 606 EPA methods) and solid-liquid extraction (506 EPA method).

The extraction methods used in this study are schematized in the Figure 2. The liquidliquid extraction has been carried out taking into account both methods 506 EPA method (use of NaCl during the process) and 606 EPA method (without NaCl). The solid-liquid extraction is related to 506 EPA method.

RESULTS Analysis by GC/MS of Standards In Figures 3, 4 and 5 are reflected the chromatograms and spectrums for DEHP, DEHA and DIDP standards, respectively.

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26.07

Figure 3. Chromatogram and MS spectrum of DEHP

23.01

Figure 4. Chromatogram and MS spectrum of DEHA.

Table 2. Identification by GC/MS for each compound. Compound DEHA DEHP DIDP

Retention time (min) 26.06 23.01 36-48

Characteristic mass (m/z) 149 129 149

A summary of the retention time and the characteristic mass used for the identification of DEHP, DEHA and DIDP is shown in Table 2.

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149

RT: 29.92 - 56.14 NL: 6.68E5

41.45

100

TIC MS 25ppm

95 90 85 80

39.64

42.33 40.73

75

43.43

70

Relative Abundance

65

43.74 44.09

60

45.19 38.94

55 37.88

50

46.17

45 36.92

40 31.18

35

33.94 35.06

37.25

47.45 47.95 48.62 49.81

35.65

55.63 50.62 51.47 52.81 54.50

30 25 20 15 10 5 0 30

32

34

36 36

38

40

42 44 Time (min)

46

48 48

50

52

54

56

Figure 5. Chromatogram and MS spectrum of DIDP

As it can be observed in Figure 5, the diisodecyl phthalate (DIDP) is not identified by only one peak in the chromatogram since it is a mixture of isomers. Therefore, for its identification it is necessary to consider the time range 36-48 min. The quantification of di(2-ethylhexyl)phthalate (DEHP) and di(2-ethylhexyl)adipate (DEHA) was based on the retention time and their characteristic mass (m/z=149 and m/z=129 for DEHP and DEHA, respectively). In the case of the DIDP the quantification was made taking into account two options: (a) Considering the most representative isomers. The quantification is based on the retention times of these isomers (39.60 min and 41.45 min) and their characteristic masses (m/z=149). (b) Considering the area of all the isomers peaks as a whole. The quantification is based on the manual integration of the total area and the characteristic mass (m/z=149). For this reason, three calibration curves were established for the DIDP: 1. Calibration DIDP t=39 : The peak at 39 min is considered as representative of this compound. 2. Calibration DIDP t=41 : The peak at 41 min is considered as representative of this compound. 3. Calibration DIDP total : The total area of all isomers is considered as representative.

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Table 3. Details of the calibration curves obtained in this study. R2 0.9983

Linear range 5-250 µg l-1

QL* 5µg l-1

0.9981

5-250 µg l-1

5 µg l-1

DIDPt=39

Calibration curve FR=6.8655*10-5+0.0004348*C (µg l-1) FR=-0.001073+0.00024809*C (µg l1 ) FR=-0.0368+0.0215*C (mg l-1)

0.9977

3.12-75 mg l-

3.1 mg l-1

DIDPt=41 DIDPtotal

FR=-0.1469+0.0446*C (mg l-1) FR=-3.4716+0.8841*C (mg l-1)

0.9994 0.9981

6.5-75 mg l-1 6.5-75 mg l-1

6.5 mg l-1 6.5 mg l-1

DEHP DEHA

1

*QL = quantification limit / FR= factor response

Deuterated Anthracene (AD10) (retention time= 16.53 min, m/z= 188) was used as an internal standard for the quantification of target compounds by GC/MS analysis. The calibration curves, the linear ranges and the quantification limits obtained for DEHP, DEHA and DIDP are shown in Table 3.

Study of the Background Pollution The background pollution was evaluated in the different stpes involved in the analytical method used to quantify the compounds of interest (DEHA, DEHP and DIDP). This problem of background contamination has been more serious in the trace analysis of plasticizers than in the studies of many other pollutants because these compounds are present in almost all equipments and reagents used in the laboratory. Different commercial reagents were evaluated (Carlo Erba, Merck…) with the aim of selecting the commercial brand most appropriate for carrying out these analyses. The obtained results are shown in Table 4. Table 4. Study of Background pollution.

1

Materials and reagents Dichloromethan e Hexane NaCl Methanol Water

Brands analysed

Brand selected

Carlo Erba1, Merck2 Carlo Erba1, Merck2, SDS3 Carlo Erba1, Merck2 SDS4 MilliQ5,Mineral water A6, Mineral water B7, mineral water C8

Merck Merck Merck SDS Mineral water A

Glass wool Teflon tube (Nitrogen)

Panreac, Carlo Erba1, Merck2 N95: Carburos metalicos (air products)

Merck ----

99% purity, analysis quality quality for organic traces 3 quality for pesticides analysis 2

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4

HPLC quality5 Ultrapure Milli Q water Mineral Water A (bottled in glass) 7 Mineral Water B (bottled in glass) 8 Mineral Water C (bottled in carton) 6

Modified liquid-liquid method

Modified solid-liquid method Sample (4ºC)

Sample (4ºC)

If it‘s necessary: Dilution with mineral water A+ NaCl

Liquid-liquid extraction 3 extractions with CH2Cl2 +1 extraction C6H6

Concentration in rotatory evaporator (2-4ml)

C6H6 addition until volume=10 ml

Drying of the extract with Na2SO4 anhidrous

Concentration with N2 (volume 1 ml)

GC/MS analysis

If necessary: Dilution with mineral water A Sample dilution with MIlliQ water (if it‘s necessary)

Solid-liquid extraction C18 cartridge

- Cartridge preparation CH2Cl2/MeOH/H2O df fffffff(Mineral water A) - Compound adsorption -Drying cartridge with N2

Elution of compounds with CH2Cl2 (3-4 ml, gravity)

C6H6 addition until volume=10 ml

Extract dried with Na2SO4 anhidrous

Concentration with N2 (volume 1 ml)

GC/MS analysis

Figure 6. Modified methods in accordance with the background pollution study.

As a consequence of this study, the analytical methodologies previously described in the experimental procedure section have been modified to avoid as much as possible the background pollution. The main modifications are related to avoid DEHA and DEHP pollution since DIDP contamination was not detected (Table 4). For this reason, mineral water A bottled in glass was used instead of MilliQ water. Furthermore, the extract concentration has been modified and the reagents most adequate have been chosen (Table 4).The modified methodologies are described in Figure 6.

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Study of the Recovery of Each Method The following analytical methodologies have been compared in order to know which one is the most appropriate for the analysis of DEHP, DEHA and DIDP:   

Modified solid-liquid extraction + GC/MS analysis Modified liquid-liquid extraction + GC/MS analysis Modified liquid-liquid extraction with addition of NaCl+ GC/MS analysis

The synthetic samples were prepared by diluting stock solution in mineral water A, obtaining concentrations of 30 µg l-1 for DEHA, 60 µg l-1 for DEHP and 1.2 mg l-1 for DIDP. The recovery results obtained for each analytical method are shown in Figure 7. As it can be observed in Figure 7, the recoveries obtained with the solid-liquid extraction were very low, probably due to polarity of the adsorbent. As a consequence, this method was rejected. In the case of liquid-liquid methodologies the recoveries were higher. It is also noticed that the NaCl addition, used to saturate water and to help the movement of the compounds to organic phase, improved the liquid-liquid method. The DEHA and DEHP recoveries were twice and four times higher respectively whereas DIDP recovery increased slightly. Therefore the liquid-liquid extraction with NaCl addition was selected for the analysis of the target pollutants.

Control of DEHA, DEHP and DIPD in Real Waters Control in different mineral waters and tap water This study was carried out with different brands of mineral waters and tap water (water B) from the city of Zaragoza (Spain).

Figure 7. % Recoveries obtained for DEHP, DEHA and DIDP using different methods of extraction.

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The main features of the studied mineral waters, taking into account that A, C, D,…are the brands, are the following:       

- Mineral water A bottled in PET. - Mineral water A bottled in glass. - Mineral water C bottled in PET. - Mineral water C bottled in Carton. - Mineral water D bottled in PET - Mineral water E bottled in PET. - Mineral water F bottled in HPDE.

The obtained results are reflected in Figure 8. It is observed that the tap water was the most polluted (11.8 µg l-1 DEHA and 29.2 µg l-1 DEHP). The mineral water C bottled in carton and the mineral water A bottled in glass presented the less concentration of DEHA (1.1 µg l-1) and DEHP (6.9 µg l-1). Considering all the studied samples, the range of DEHA concentration was 1.1-3.5 µg l-1 and with respect to DEHP the range was 6.9-23.3 µg l-1. The DIDP compound was not detected in the analyzed waters.

Figure 8. Control of DEHP, DEHA and DIDP in tap and mineral waters.

Furthermore, from Figure 8 it can be concluded that the concentration of the target pollutants depends on the material of which the bottle is made of. In order to compare in a better way the concentrations of the target compounds, in Figure 9 are shown the results for mineral water A and C, bottled in glass, PET and carton. The main difference between mineral waters was due to the DEHP concentration, which was lower in glass bottle.

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Figure 9. Concentration of DEHP and DEHA in mineral water A and C.

Figure 10. Concentration of DEHP and DEHA in mineral waters bottled in PET.

In Figure 10, the mineral waters bottled in PET are compared. As it can be observed, the DEHA concentration was very similar for all the analyzed mineral waters. However, the DEHP concentration showed a great variation. In fact, the DEHP concentration in mineral waters A and D was twice the concentration in mineral water C and E.

Control in wastewaters generated in PVC manufacturing and surrounding groundwater In this section, it is shown the control of the target compounds in both real wastewaters produced during the PVC manufacturing and in the groundwater used by the manufacturing plant for their own supply. The different analyzed wastewaters were the following:

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Wastewater I: Generated during the process of polymer synthesis. Wastewater II: Produced during the process of polymer transformation by the additives addition (plasticizers). This transformation takes place at high temperature and the analyzed wastewater was the refrigeration water used in this installation.

Figure 11. Presence of DEHP and DEHA in wastewaters and groundwater

Figure 12. Chromatogram of groundwater.

Two different samples of these wastewaters were analyzed and the results obtained are shown in Figure 11 and it is observed that the DEHP concentration was always higher than DEHA concentration. DIDP compound was not detected in any sample. The chromatograms obtained for analyses of groundwater and wastewater II (sampling b) are reflected in Figure 12 and 13, respectively.

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The chromatogram of wastewater II (sample b) (Figure 13), shows previous peaks to DEHP peak (m/z=149) which are probably related to some plasticizers containing a lower number of carbon atoms. These substances appear in significant concentrations.

Figure 13. Chromatogram of wastewater II (sample b).

CONCLUSION -

-

-

-

The study of background pollution during the determination of di(2-ethylhexyl) phthalate, di(2-ethylhexyl) adipate and diisodecyl phthalate in aqueous samples is essential and the use of solvent of adequate quality is important as well. Due to the presence of these plasticizers in almost all equipments and reagents used in the laboratory, their control in each stage of the analytical methodology is necessary. The liquid-liquid extraction with NaCl addition followed by GC/MS was selected for the analysis of the target pollutants since the highest recoveries were obtained (80.5% for DEHA, 119% for DEHP and 139.9 % for DIPD). The control of the target pollutants in groundwater, wastewaters and mineral waters indicate that the presence of DEHP was important in all the samples and its concentration was higher that the DEHA concentration. DIDP compound was not detected in any sample. The detected concentration of DEHP and DEHA is significantly higher in the case of tap water than in the case of the analyzed mineral waters. In all the samples, the DEHA concentration always presented a concentration lower than the limit established by the EPA (0.4 mg/l) for drinking waters [17]. Respect to the control of these compounds in groundwater and wastewaters, it was observed that the DEHP concentration is always higher than the DEHA concentration. The wastewater generated during the process of polymer transformation by the additives addition is the most polluted by these plasticizers.

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REFERENCES [1] [2] [3] [4] [5] [6] [7] [8] [9] [10] [11] [12] [13] [14] [15] [16] [17]

Fromme, H., Kuchler, T., Otto, T., Pilz, ,K. Muller, J. & Wenzel, A. (2002). Occurrence of phthalates and bisphenol A and F in the environment. Water Research, 36, 14291438. Staples, C. A., Peterson, D. R., Parketon, T. F. & Adams, W. J. (1997). The environmental fate of phthalate esters: a literature review. Chemosphere, 35, 667-749. Harris, C. A., Henttu, P., Parker, M. G. & Sumpter, J. P. (1997). The estrogenic activity of phthalate esters in vitro. Environmental Health Perspectives, 105, 802-811. Jobling, S., Reynolds, T., White, R., Parker, M. G. & Sumpter, J. P. (1995). A variety of environmentally persistent chemical, including some phthalate plasticizers, are weakly estrogenic. Environmental Health Perspectives, 103, 582-587. Combes, R. D. (2000). Endocrine disruptors: a critical review of in vitro and in vivo testing strategies for assessing their toxic hazard to humans. ATLA, 28, 81-118. Birkett, J. W. & Lester, J. N. (2003). Endocrine disrupters in wastewater and sludge treatment processes. London, Lewis Publisher. United Kingdom Economics Account (2000). Endocrine disrupting substances in the environment: The environment Agency‘s strategy. Environment Agency. U.S. Environmental Protection Agency. (1997). Special Report on Environmental Endocrine Disruption: An effects Assessment and Analysis, Report Nº EPA/630/R96/012, Washington D.C. Oslo and Paris Commisions. (1998). OSPAR strategy with regard to hazardous substances. OSPAR convention for the protection of the marine environment of the North-East Atlantic. OSPAR 98/14/1, Annex 34. U.S Environmental Protection Agency. (1991). National Primary Drinking Water Regulations; Fed. Reg., Part 12, 40 CFR Part 141, p.395, U.S. Environmental protection agency, Washington, DC. Decision nº 2455/2001/EC of the European Parliament and of the council of 20 November 2001 establishing the list of priority substances in the field of water policy and amending Directive 2000/60/EC. World Health Organization Guidelines for Drinking Water quality (1993), set up in Geneve. International Agency for Research on Cancer. (1982). Some industrial chemicals and dyestuffs. International agency for Research on Cancer, IARC Monographs on the Evaluation of the Carcinogenic Risk of Chemicals to Humans, 29, 257-267. Guidelines for drinking water quality (1996). 2nd ed. Vol. .2. Health criteria and other supporting information. World Health Organization, Geneva. U.S. EPA. (1995). 600/R-95-131. Method 506. Determination of Phtalate and adipate esters in dringing water by liquid-liquid extraction or liquid-solid extraction and chromatography with photoionization detection. CD-1DWSU. U.S. EPA. (1996). 821/B-96-005, Method 606., Determination of phtalates in municipal and industrial wastewater, 40 CFR, part 136 Appendix A, CD 40_136. U.S. EPA (1997). Public Health Goal for Di(2-Ethyl-hexyl)Phthalate (DEHP) in drinking water, pesticide and environmental Toxicology section, Office of Environmental Health Hazard Assessment, California Environmental Protection Agency.

INDEX A abatement, 42, 108 absorption, 9, 87, 88, 93, 95, 98, 122, 136, 137, 181, 183, 185, 195 acceleration, 119, 252 accelerator, 95 acceptor, 53, 56, 57, 59, 137, 140, 162 accounting, 280 accuracy, 112, 123 acetate, 54, 55, 56, 57, 63, 68, 69, 70, 140, 141, 149, 152, 170, 182, 237 acetic acid, 53, 68, 69, 140, 157, 201, 262 acetone, 61, 69, 201 acetonitrile, 226 acid, 10, 13, 17, 20, 45, 53, 55, 57, 62, 63, 68, 69, 75, 76, 78, 79, 84, 87, 136, 140, 147, 149, 150, 151, 157, 177, 180, 184, 185, 186, 191, 192, 193, 195, 201, 202, 212, 219, 226, 237, 262, 277, 297, 330, 331 acidic, x, 6, 9, 17, 30, 36, 50, 56, 60, 62, 80, 81, 83, 85, 88, 106, 136, 141, 181, 182, 206, 236 acidification, 67 acidity, 281 ACM, 139 activated carbon, 82, 108, 320 activation, xii, xvi, 113, 158, 197, 320 activation energy, 158 active site, 324 acute, 114, 121, 130, 154, 230, 276 adaptation, 64, 75, 158, 233, 246 additives, 2, 22, 108, 114, 344, 346 adhesion, 7 adipate, xvi, 333, 334, 335, 338, 346, 347 adjustment, 17, 84, 236, 322 administration, 126 adsorption, xiv, xvi, 2, 5, 7, 31, 50, 62, 83, 88, 89, 95, 96, 108, 113, 114, 115, 181, 183, 185, 187,

188, 191, 192, 249, 250, 254, 255, 259, 260, 261, 262, 281, 319, 321, 322, 323, 324, 325, 326, 327, 329, 330, 331 aerobic, xi, 51, 52, 58, 65, 73, 115, 135, 143, 144, 147, 149, 152, 156, 160, 162, 163, 168, 170, 172, 185, 186, 191, 192, 238, 240, 267, 268, 271, 272, 273, 276 aerobic bacteria, 58, 185 aerosols, 136 Africa, 46, 47, 280, 282, 297, 300, 306, 308, 317 Ag, 330 agar, 57, 59, 182, 311, 321 age, 12, 191 agents, xiii, 10, 46, 60, 61, 82, 83, 114, 200, 215, 216, 217, 219, 221, 222, 225, 226, 227, 231, 238, 240, 297, 329 aggregates, 7, 83, 87, 239 aggregation, 5, 7, 240, 243 aging, 5 agricultural, xiii, xiv, xv, 30, 49, 199, 213, 230, 235, 236, 237, 238, 240, 242, 244, 245, 246, 279, 291, 304, 310, 311, 312, 314, 316 agricultural crop, xiv, 236, 244 agriculture, ix, xiii, 51, 52, 65, 230, 235, 246, 288, 291, 303, 304, 309, 311, 312, 313, 314, 316, 317 air, xiii, 14, 31, 84, 86, 87, 89, 107, 114, 127, 136, 143, 147, 160, 178, 203, 235, 245, 255, 262, 339 air quality, 136 alcohol, 62, 68, 69, 180, 181, 190, 193, 202, 273 alcohols, 53, 54, 59, 61, 62, 141, 237 aldehydes, 61, 62 algae, 122, 144, 154, 320 Algeria, 300 aliphatic compounds, 54, 201 ALK, 54 alkali, 61, 62, 85, 92, 95, 108, 109, 139 alkaline, 6, 17, 30, 50, 62, 98, 106, 109, 137, 139, 146, 148, 180, 204, 245 alkalinity, 16, 156, 161, 162, 283, 291, 292

346

Index

alkane, 79 alkanes, 54, 75, 76 alternative, ix, xi, xii, xiii, xvi, 1, 42, 82, 122, 136, 143, 147, 149, 153, 172, 173, 175, 199, 215, 221, 228, 235, 237, 244, 271, 309, 319, 320, 329 alternatives, xi, 30, 82, 135, 148, 163, 230, 243, 282 aluminosilicates, 92, 102, 103, 104 aluminum, ix, 1, 5, 7, 9, 10, 13, 15, 17, 18, 19, 23, 25, 27, 28, 41, 43, 45, 46, 84, 108, 109 Aluminum, 6, 32, 43, 45, 46 aluminum surface, 10 ambient air, 136 amendments, 77 AMF, 241 amide, xvi, 320, 329, 330 amine, 330 amino, 53, 54, 75, 180, 186, 195, 236, 329 ammonia, 154, 155, 156, 157, 158, 159, 160, 162, 163, 164, 166, 170, 173, 175, 177, 239 ammonium, 23, 39, 44, 50, 114, 154, 170, 177, 178 amorphous, x, 7, 81, 83, 90, 92, 98, 106, 108 amorphous phases, x, 81, 92, 98, 106 Amsterdam, 132, 178, 234 anaerobe, 139 anaerobes, 52, 53, 58 anaerobic bacteria, 58, 74, 76, 138, 139, 140, 141, 185, 194, 195, 237, 240 anaerobic digesters, 156, 158, 236, 245 anaerobic granular sludge, 78 anaerobic sludge, 77, 173, 178 Angola, 300 animal health, xiii, 235 animals, xvi, 60, 238, 319, 320 anode, ix, 1, 5, 8, 9, 10, 11, 16, 17, 19, 20, 23, 47, 287, 288 anodes, 6, 7, 10, 11, 17, 22, 44 anoxic, 75, 79, 141, 155, 156, 160, 163, 170, 171, 178, 186 antagonists, 56 antenna, 75 anthropogenic, ix, 49, 50, 51, 52, 53, 63, 66, 74, 122, 136, 154, 302 antibiotic, 222, 242, 244, 246, 277 antibiotic resistance, 222, 242, 246 antibiotics, 180, 233, 242, 246 antimony, 65 antioxidant, 199 aquaculture, xii, 153, 175 aqueous solution, xvi, 11, 23, 107, 148, 319, 320, 330, 331 aqueous solutions, xvi, 11, 23, 148, 320, 330, 331 aqueous suspension, 46 aquifers, 77, 78

Arabia, 1, 279, 300 arbuscular mycorrhizal fungi, 241 archaea, 56, 59 argument, 315 arid, 244, 245, 246, 315 Arizona, 241, 247 aromatic compounds, 54, 113, 268 aromatic hydrocarbons, 53, 59, 62, 77, 78, 113, 114, 115, 124 aromatics, 201 arsenic, 12, 65, 108, 114, 122, 237, 262 arsenobetaine, 122 ash, x, xv, 2, 12, 17, 22, 23, 24, 25, 28, 30, 31, 32, 33, 34, 35, 36, 37, 38, 39, 40, 41, 42, 43, 44, 46, 82, 108, 109, 110, 237, 251, 279, 281, 282, 284, 285, 286, 287, 288, 289, 290, 292, 296 Asia, 180, 193 Asian, 316 Aspergillus niger, 184, 189, 194 assessment, 108, 109, 112, 124, 126, 130, 244, 275, 276, 278, 283 assessment techniques, 112 assimilation, 155, 183 assumptions, 315 ASTM, 282, 296 asymptotically, 118 Athens, 131 Atlantic, 77, 347 atmosphere, 66, 89, 112, 136, 154, 202, 237 atmospheric pressure, 61, 136 atoms, 345 ATP, 55, 56 attachment, 321 AU, 22 Australia, 113, 120, 121, 125 Austria, 167, 215, 216, 217, 222, 230, 233, 280 autarky, 177 automation, 224 autotrophic, xii, 72, 76, 137, 145, 146, 153, 155, 158, 160, 161, 170, 171, 173, 174, 175, 176, 177, 178, 202 availability, 2, 104, 115, 123, 142, 167, 168, 284, 300, 303

B Bacillus, 185, 194, 238 back, xv, 13, 36, 237, 244, 253, 280, 301 bacterial, xiii, 76, 77, 121, 152, 182, 194, 216, 236, 238, 241, 242, 313, 315 bacterial strains, 182, 242 bacterium, 75, 76, 77, 79, 143, 145, 148, 151, 242 ballast, 60

Index bananas, 303 band gap, 11 Bangladesh, 308 barium, 114 barrier, 31, 76, 222 batteries, 66 battery, xvi, 319, 320 beet molasses, 48, 193, 194 behavior, 20, 83, 107 Beijing, 86 Belgium, 48 benefits, x, xiv, 2, 167, 236, 237, 244, 246, 278, 308, 310, 312 benign, xi, 135 benzene, 54, 58, 76, 80, 121, 126 beverages, 60 bicarbonate, 170 binding, x, xvi, 82, 84, 88, 101, 106, 320, 329, 330 bioaccumulation, 113, 121 bioaerosols, 238 bioassays, 122 bioavailability, 114, 126 biocatalyst, 151, 177 bioconversion, 143, 149 biodegradability, 114, 216, 266, 269, 273, 276, 277 biodegradable, 50, 59, 201, 219, 221, 266, 267, 268, 269, 270, 271, 276, 278 biodegradation, xiv, 51, 52, 54, 59, 60, 63, 67, 68, 69, 70, 71, 74, 76, 80, 112, 113, 114, 115, 193, 194, 212, 219, 237, 265, 269, 273, 275, 276, 278 biodiversity, 304 biofilms, 63, 140, 152 biogas, 69, 168, 169, 170, 177, 237, 240, 244, 246 biological activity, 146 biological processes, 2, 114, 137, 152, 154, 243, 271 biological systems, 137 biomarker, 122, 241 biomass, xvi, 51, 64, 139, 143, 144, 146, 154, 158, 160, 162, 164, 166, 173, 174, 180, 193, 203, 206, 208, 209, 210, 211, 239, 241, 244, 266, 268, 269, 315, 319, 320, 324, 325, 326, 328, 329, 330, 331 biomaterials, 320 bioreactor, 52, 64, 73, 74, 78, 137, 141, 143, 145, 146, 150, 151, 174, 190, 193, 320, 321, 331 Bioreactor, 150, 173, 174 bioreactors, 53, 63, 64, 69, 137, 139, 140, 142, 145, 146, 173 bioremediation, 54, 67, 72, 76, 78, 79, 194 biosorption, 320, 324, 326, 328, 329, 330 biosphere, 49, 154 biota, 122, 125, 126, 247 biotechnological, xi, 135, 136, 142, 147, 148 biotechnology, xi, 51, 52, 71, 72, 135

347

biotic, 146 biotransformation, 51, 52, 63, 65, 78, 80 bipolar, 26, 46, 47, 217, 218, 228, 231 bisphenol, 347 bleaching, 8, 12 blends, 107, 278 blocks, 142 blood, 154, 276 blue baby, 154 body weight, 230 boilers, 30 boiling, 61 bonding, 7, 92, 95, 329 bonds, x, 12, 82, 102, 104, 216, 236 boreholes, 302 Boron, viii, 10, 215, 218 Boston, 149 Botswana, 300 Brazil, 131, 280 breakdown, 220, 225 breeding, 60 Britain, 8, 296 British Columbia, 151 broad spectrum, ix broodstock, 175 Brussels, 245 bubble, 15, 16, 17, 18, 20, 139, 148 bubbles, 7, 8, 11, 13 Bubbles, 45 buffer, 17, 170, 182, 184 Bureau of Reclamation, 311 burning, 89, 91 butadiene, 61 butyric, 69 by-products, 12, 50, 68, 212, 269

C Ca2+, x, 31, 81, 100, 106, 285 cadmium, x, 22, 44, 65, 82, 102, 114, 237, 244, 331 calcium, 10, 11, 30, 43, 52, 63, 77, 92, 114, 225, 234 calcium carbonate, 10, 30, 77 calibration, 123, 322, 338, 339 Canada, 43, 47, 77, 124, 297, 311 cane sugar, 196 capillary, 88 capital cost, 13, 261 carbohydrates, 53, 54, 68, 144, 206, 236 Carbon, 45, 149 carbon atoms, 345 carbon dioxide, 55, 68, 69, 72, 74, 137, 144, 152, 237, 267 carbon monoxide, 150

348

Index

carbonates, 30, 52, 63, 114 carboxyl, 329, 330 carboxylates, 53 carboxylic acids, 201 carcinogenic, 154, 230 carrier, 64, 156, 192, 193, 194 case study, xi, 46, 111, 113, 124, 127, 130, 276, 278 casein, 58, 60 cash crops, 280 catalysis, 67 catalyst, xii, 109, 136, 197, 202, 203, 206, 207, 208, 211, 223 catalytic properties, 287 catchments, 303, 304 catechol, 54, 79, 202 cathode, 5, 8, 10, 11, 16, 17, 23, 287, 288 cation, 7, 9, 15, 106, 107 cattle, 199 cell, 13, 15, 23, 57, 67, 69, 72, 113, 137, 142, 147, 154, 185, 187, 188, 192, 203, 209, 211, 231, 240, 242, 284, 288, 294, 295, 330 cell death, 209 cell membranes, 113 cell metabolism, 240 cell organelles, 188 cell surface, 185, 187, 188, 211 cellulose, 184, 190, 193, 194, 281, 329 cement, 82 ceramics, 64, 82, 108, 109, 110, 216, 218 CH3COOH, 56, 68 CH4, xi, 70, 111, 113 channels, 250, 281 charged particle, 2 cheese, 60, 77 chelating agents, 221, 225 chemical approach, 130 chemical bonds, x, 82, 104, 236 chemical engineering, 15 chemical industry, 199 chemical oxidation, xii, 197, 202, 204, 206, 207, 211, 219, 221, 269, 320 chemical properties, 241, 242, 243, 246, 247 chemical reactions, 106, 114 chemical reactivity, 100 chemical stability, xiii, 95, 215, 216, 217, 220 chemical vapor deposition, 216 chemicals, 11, 13, 121, 124, 156, 219, 222, 224, 233, 239, 266, 268, 271, 272, 273, 314, 334, 347 chicken, 272, 273 chickens, 277 China, 46, 81, 84, 86, 87, 316 chitin, 329

chloride, xii, 23, 28, 107, 197, 202, 207, 221, 225, 311, 313, 334 chlorination, 154 chlorine, 12, 18, 19, 221 chlorobenzene, 201 chlorophyll, 144 Cholera, 305 chromatograms, 337, 345 chromatography, xvi, 226, 333, 347 chromium, 9, 16, 17, 18, 19, 20, 21, 67, 109, 114, 262, 330 Chromium, 16, 19, 21, 32, 43, 44, 45, 47 circulation, 63, 115, 127, 203, 230 civil engineering, 94 clams, 144 classes, 183, 242, 334 classical, 51 classification, 119, 120, 123, 190 clay, 64, 82, 85, 106, 109, 244 cleaning, xv, 10, 60, 61, 148, 201, 202, 237, 280, 296 cloning, 241 CO2, xii, 17, 39, 49, 53, 55, 56, 69, 70, 85, 96, 137, 140, 144, 149, 153, 168, 173, 202, 219, 220 coagulation, ix, 1, 2, 5, 7, 12, 13, 23, 46, 47, 48, 50, 83, 84, 281, 284, 289, 296, 297 coagulation process, 289 coal, 18, 107, 136, 280 coastal areas, x, 111, 112, 120, 124 coatings, 216 cobalt, 65 coconut, 251, 262 co-existence, 70 coffee, ix, xv, 2, 11, 25, 36, 37, 41, 279, 280, 281, 282, 283, 284, 285, 286, 287, 288, 289, 290, 291, 292, 293, 296, 297, 303 coke, 136 collagen, 190 colloidal particles, 2, 4, 5, 7, 13, 142 colloids, 2, 3, 31 Columbia, 151 combined effect, 288 combustion, 30, 32, 107, 108, 110, 136, 180 combustion processes, 136 commodity, 250, 280 communities, xiii, 57, 58, 74, 79, 121, 235, 241, 242, 243, 245, 246, 250, 261, 303, 304, 306, 317 community, xiv, 43, 59, 80, 115, 122, 126, 236, 239, 241, 242, 243, 245, 246, 255, 303 competition, 70, 140, 141, 301 competitive advantage, 145 complement, 122, 124, 126 complexity, ix, 1, 14, 16, 122, 123 compliance, 266

Index components, x, 50, 59, 70, 81, 83, 85, 87, 88, 95, 104, 106, 122, 180, 188, 194, 231, 238, 269, 271 composites, 108 composition, xiv, 16, 30, 31, 32, 36, 46, 50, 51, 60, 61, 64, 67, 91, 94, 106, 112, 113, 120, 122, 177, 199, 203, 204, 211, 234, 236, 245, 246, 312, 320 compost, 146, 244 composting, 199, 212, 238 compressive strength, 83, 84, 88, 94, 95, 96, 98, 100, 106 computation, 323 computer simulations, 315 concrete, 109, 233, 281 condensation, 23, 180 conditioning, 12, 170, 237 conductive, 11, 14, 216, 217, 218 conductivity, ix, 1, 12, 14, 22, 23, 27, 29, 31, 32, 43, 127, 202, 203, 206, 207, 208, 219, 226, 228, 232, 233, 239, 282, 283, 286, 288, 291, 292, 311, 313, 315, 316 configuration, 11, 36, 38, 156, 157, 163, 166, 169, 171, 173, 284 conflict, 300, 302, 303, 316 Congress, 194 conservation, 83, 128, 202, 315 construction, x, xiii, 82, 83, 199, 215, 217, 222, 255, 257, 259, 261 construction materials, 83 consultants, 130 consumers, xiv, 230, 236 consumption, x, xii, xv, 2, 9, 12, 13, 17, 20, 23, 24, 25, 26, 27, 28, 36, 41, 43, 44, 59, 83, 102, 144, 153, 169, 202, 207, 211, 230, 279, 282, 284, 287, 288, 289, 291, 293, 297, 300, 306, 310, 315, 317 contact time, xvi, 139, 227, 231, 319, 321, 324, 325, 327, 328, 329 contaminant, 121, 126, 210, 211 contaminants, ix, xi, xiii, 1, 5, 21, 109, 111, 114, 120, 122, 126, 216, 219, 243, 250, 334 contaminated soils, 109 contamination, 82, 173, 230, 244, 304, 334, 339, 341 control, xvi, 10, 14, 79, 108, 136, 141, 145, 148, 149, 150, 151, 152, 156, 158, 163, 164, 194, 210, 211, 229, 239, 241, 242, 271, 276, 288, 296, 333, 334, 344, 346 conversion, 73, 77, 83, 137, 138, 140, 142, 143, 144, 147, 149, 154, 180, 236, 293 conversion rate, 140 cooling, 11, 60, 93, 225 cooling process, 93 Copenhagen, 142, 144, 152 copper, x, 22, 44, 65, 79, 82, 102, 114, 237, 330, 331 copper oxide, x, 82, 102

349

corn, xvi, 229, 276, 319, 320, 321, 330, 331 corona, 224 correlation, xvi, 242, 320, 326, 328, 330 correlation coefficient, xvi, 320, 326, 328, 330 correlations, 269 corrosion, 10, 11, 23, 45, 114, 122, 140, 170 corrosive, 142, 237 corruption, 306, 316 cost benefits, 312 cost of power, 296 cost saving, 164, 313, 314, 315 cost-benefit analysis, xv, 309 cost-effective, ix, 2, 130, 320, 329 costs, xii, xiii, xv, 25, 30, 50, 51, 136, 153, 156, 158, 163, 164, 165, 168, 169, 172, 175, 199, 221, 227, 233, 235, 250, 296, 308, 309, 313, 314, 317, 321 cotton, 269, 273, 330 coupling, 137, 143 crack, 92 cracking, 61, 62, 88, 217 crops, 238, 239, 280, 303 crude oil, 53, 54, 61, 62, 73, 79, 80, 151 crustaceans, 122 crystalline, x, 81, 90, 91, 92, 93, 96, 97, 98, 100, 102, 103, 104, 106 crystalline solids, 103 crystallinity, 91, 92 crystallization, x, 81, 84, 90, 98, 102, 103 crystals, 91, 92, 98, 101, 102, 106 cultivation, 149, 154, 182, 188, 190, 211, 229, 242, 280, 308, 320 culture, 51, 52, 58, 59, 63, 65, 70, 137, 148, 160, 182, 184, 185, 192, 203, 210, 211, 213, 267, 321 CVD, 216, 218 cyanobacteria, 144, 186 cyanobacterium, 194 cycles, 161, 163 cycling, xiii, 235 Cyprus, 277, 278 cysteine, 57, 58 cysts, 9 cytochrome, 56 cytometry, 147, 246 cytoplasm, 185, 188, 192

D dairy, 50, 53, 58, 60, 63, 75, 80, 143, 151, 277 dairy industry, 50, 60, 80 data set, 128 database, 87, 92, 102 death, 154, 209 decane, 54

350

Index

decay, 116 decision makers, 126, 130 decisions, 125, 130, 275 decomposition, xiii, 54, 55, 59, 68, 69, 70, 74, 85, 98, 109, 154, 183, 184, 186, 188, 212, 235, 241 decomposition reactions, 98 defects, 113 deficiency, 13, 206, 211, 267 deficit, 206 definition, 59 deforestation, xv, 299, 303, 307 degradation, xiii, xv, 12, 50, 75, 76, 78, 79, 80, 114, 115, 121, 124, 145, 146, 181, 183, 185, 191, 198, 201, 202, 203, 204, 207, 217, 225, 228, 230, 231, 235, 236, 237, 240, 241, 242, 283, 299, 304, 307 degradation mechanism, 185, 191, 202 degradation process, 114, 121, 236 degrading, 75, 141, 227 degree of crystallinity, 92 dehydrogenase, 243 demobilization, 67 denitrification, xi, xii, 153, 154, 155, 156, 157, 159, 160, 161, 163, 164, 165, 167, 168, 170, 171, 173, 175, 176, 178 denitrifying, 67, 143, 151, 155, 156, 158, 161, 168, 173, 174, 175 density, ix, xv, 1, 14, 15, 18, 19, 20, 21, 23, 24, 26, 27, 36, 42, 65, 83, 87, 88, 89, 94, 95, 96, 115, 117, 118, 119, 126, 128, 129, 181, 182, 218, 221, 223, 224, 227, 228, 229, 231, 232, 240, 279, 284, 289, 294, 295 Department of Energy, 249 Department of Interior, 132 deposition, 216, 218 deposits, x, 10, 47, 50, 52, 53, 63, 64, 65, 66, 67, 111, 112, 297 depreciation, 25 derivatives, 12, 75, 202 desalination, xvi, 309, 312, 314, 315, 316 desert, 317 destruction, 83, 107, 154, 200, 201, 307 detection, 140, 242, 244, 347 detention, 27, 33, 35, 288 detergents, 46 detoxification, 194, 195, 277 developed countries, 82 developing countries, xv, xvi, 82, 299, 306, 307, 308, 319, 320 deviation, 203 dextrose, 188, 321 dialysis, 310, 311 diamond, ix, xiii, 10, 11, 42, 215, 216, 217, 220, 222, 224, 228, 229, 232, 233

Diamond, viii, 11, 45, 46, 215, 216, 218, 219, 228, 233, 234 diamonds, 10, 222 diaphragm, 223 dielectric constant, 3 diesel, 61 differentiation, 293 diffraction, 30, 83, 88, 92 diffusion, 21, 58, 106, 173 digestion, 78, 87, 122, 141, 168, 171, 172, 173, 193, 199, 236, 238, 239, 240, 244, 253 dioxin, 201 direct costs, 25 discharges, 112, 118, 121, 124, 125, 126, 127, 130, 131, 132, 266, 271, 272, 273, 310, 313 diseases, 243, 303, 305, 307 disinfection, xiii, 215, 217, 221, 231, 232, 233 disorder, 154 dispersion, xi, 4, 5, 85, 111, 112, 114, 122, 123, 124, 125, 126, 127, 128, 130 dissociation, 30 dissolved oxygen, xi, 55, 135, 144, 145, 147, 150, 157, 158, 159, 162, 166, 189, 267, 269 distillation, 61, 62, 179, 314 distilled water, 32, 321 distribution, xii, 4, 53, 112, 113, 114, 121, 143, 151, 164, 179, 180, 181, 186, 187, 216, 240 diversity, xiv, 236, 241, 243 division, 210 DNA, 57, 245, 246 donor, 56, 57, 69, 139, 140, 142, 155, 158, 173, 174 donors, 53, 55, 57, 60, 144, 149, 152, 161 doped, ix, xiii, 10, 42, 215, 216, 217, 220, 222, 233 doping, 11 dosage, xvi, 20, 30, 319, 325, 326, 327 dosing, 15, 18 double bonds, 12, 216 drainage, 76, 140, 151, 303 dream, 307 drinking, x, xiii, xvi, 12, 81, 84, 107, 108, 148, 154, 175, 177, 215, 231, 233, 306, 308, 333, 334, 335, 346, 347 drinking water, xiii, xvi, 12, 107, 108, 148, 154, 175, 177, 215, 231, 233, 306, 308, 333, 334, 335, 346, 347 drought, 310 drugs, 221, 233 dry matter, 205, 239 drying, 108, 251, 254, 259, 261, 304 durability, 84, 98, 102 duration, 52, 71, 87, 229 dust, 251 dyes, 8, 320, 330

Index

E E. coli, 240 earth, xiii, 61, 65, 98, 109, 235, 250 ecological, 51, 112, 120, 121, 124, 126, 306 ecology, 77, 308 economic disadvantage, 143 economics, 142 ecosystem, xiii, xiv, 120, 126, 215, 231, 236, 242, 243, 244, 246, 300 ecosystems, 112, 114, 125, 130, 222, 241 ECSC, 245 effluents, xiv, xvi, 12, 43, 46, 49, 50, 66, 67, 122, 156, 157, 158, 162, 170, 173, 178, 199, 202, 212, 231, 265, 266, 267, 268, 269, 273, 277, 278, 282, 296, 297, 333, 335 Egypt, 300, 301, 307 elaboration, xii, 197, 198 electric charge, 2 electric conductivity, 202, 203, 207 electric field, 13 electric power, 22, 41, 42 electrical conductivity, 12, 27, 31, 32, 43, 206, 208, 282, 283, 286, 291, 292 electrical power, 17, 136 electrical properties, 2, 11 electricity, 250, 255, 261 electrochemical measurements, 45 electrochemical reaction, 8, 154 electrochemistry, 47 electrodeposition, 45 electrodes, ix, xiii, xv, 1, 5, 8, 9, 10, 11, 13, 15, 16, 17, 18, 20, 21, 22, 23, 25, 26, 27, 40, 41, 42, 43, 45, 46, 215, 216, 217, 218, 219, 220, 222, 223, 224, 228, 229, 230, 231, 232, 233, 279, 282, 284, 288, 294, 295, 297 electroflotation, 7, 11, 14, 16 electrolysis, xv, 7, 11, 13, 17, 18, 19, 24, 33, 220, 221, 279, 284, 288 electrolyte, x, 2, 9, 16, 22, 23, 24, 25, 31, 36, 41, 42, 43, 48, 286, 287, 288, 289, 292 electrolytes, x, 2, 11, 12, 22, 25, 37, 42, 45, 284, 288 electromagnetic, 61 electron, 11, 13, 53, 54, 55, 56, 57, 59, 60, 69, 72, 77, 87, 92, 137, 139, 140, 142, 144, 149, 155, 158, 161, 162, 173, 174, 188, 266, 267, 268, 269 electrons, 9, 22, 41, 53, 55, 72, 92, 106, 163, 267 electrophoresis, 7 electroplating, 11, 66, 109, 320 electroreduction, 297 electrostatic force, 3, 4 e-mail, 49, 265 emission, 49, 66, 108, 136, 177, 313

351

employment, 280 emulsification, 61 emulsions, xiii, 61, 62, 215 endocrine, 222, 230, 233, 334 endocrine system, 334 endothermic, 89, 90, 91, 96, 97, 98 energy consumption, x, xii, 2, 13, 17, 25, 27, 28, 44, 153, 169, 297 energy efficiency, 11, 220 energy supply, 136 England, 308 environmental conditions, 114, 125, 126, 147, 208 environmental effects, 124, 136, 306 environmental factors, 112 environmental impact, 125, 130, 175 environmental protection, 136 Environmental Protection Agency, 130, 133, 147, 334, 347 environmental regulations, 8, 82, 281 environmental resources, xi, 111 enzymatic, 67, 180, 242, 243 enzymes, 54, 60, 63, 68, 141, 183, 186, 192, 244 EPA, 130, 131, 133, 147, 336, 346, 347 epithelia, 122 epoxy, 61 epoxy resins, 61 equilibrium, 50, 104, 116, 315, 323, 324, 326, 329 equilibrium state, 104 erosion, 65, 303 Escherichia coli, 231, 232, 240, 245, 313 esters, 61, 226, 347 estuarine, 151 ethanol, 54, 55, 56, 57, 69, 70, 173 Ethanol, 68 Ethiopia, 280, 297, 300, 301 ethyl alcohol, 180 ethylbenzene, 54, 78 ethylene, 61 Eulerian, 123 Euro, 165, 314 Europe, 216, 231, 245, 280, 300 European Parliament, 233, 347 European Union, 334 eutrophication, 304 evaporation, xii, 112, 114, 197, 199, 226, 320 evolution, 9, 17, 125 examinations, 232 excess supply, 239 excrements, 221 excretion, 306 exothermic peaks, 91, 96 Expert System, 123 expertise, xiv, 236

352

Index

exploitation, xiii, 56, 66, 112, 122, 147, 244, 302 exports, 280 exposure, 124, 290, 307, 331, 334 Exposure, 113 extraction, x, xi, xvi, 50, 66, 87, 109, 111, 112, 113, 124, 125, 127, 130, 198, 205, 206, 226, 320, 333, 334, 335, 336, 341, 342, 346, 347

F fabric, 269, 273 failure, 156, 170, 306, 310 family, 72, 280 FAO, 293, 311, 312, 313, 315, 316 farmers, 240 farming, 63, 240 farmland, 240 farms, 49, 305 fat, 12, 60, 210, 228 fats, 60, 209, 236, 280 fatty acids, 55, 60, 62, 75, 76, 141, 236 February, 198, 277, 278, 316, 317 feces, 236, 240 Federal Register, 133 feeding, 146, 190, 191, 192, 231 feedstock, 142 feldspars, x, 81, 92, 97, 98, 102, 106 Fenton‘s reagent, 199, 200, 202 fermentation, xii, 54, 67, 68, 69, 70, 179, 180, 182, 190, 191, 193, 194, 237, 281 fern, 331 ferrous ion, 200, 201 fertility, 238, 243, 244, 303 fertilization, 239, 243 fertilizer, xiii, 66, 107, 142, 154, 172, 199, 235, 237, 238, 239, 240, 241, 242, 243, 245, 284 fertilizers, xiii, 50, 62, 63, 66, 235, 238, 239, 241, 244, 246, 305, 306 fiber, 12, 78, 278 field trials, 240 fillers, 64 film, 77, 192, 217, 313 filter feeders, 115 filters, 18, 84, 139, 182, 208, 310, 311 filtration, ix, 2, 5, 13, 14, 46, 74, 148, 174, 187, 193, 203, 204, 206, 208, 229, 311, 313, 317 financial support, 330 Finland, 80, 296 fish, 121, 122, 154, 156, 175, 305 fish production, 175 fishing, 304 fixation, 75, 102, 154, 202 flame, 283

flavonoids, 334 flexibility, 122, 170, 334 float, 312 flocculation, ix, 1, 2, 5, 7, 22, 31, 84, 115 flooding, 9, 303 flotation, 7, 11, 14, 15, 18, 20, 78, 107, 204 flow, xiv, 11, 15, 16, 22, 33, 35, 41, 51, 59, 64, 82, 120, 128, 137, 140, 147, 162, 169, 217, 222, 223, 224, 225, 227, 228, 229, 230, 232, 246, 249, 251, 252, 253, 255, 256, 257, 258, 259, 287, 288, 293, 294, 301, 304, 310, 311, 323 fluctuations, 166, 280 flue gas, xi, 135, 136, 137, 138, 139, 142, 147, 149, 237 fluid, ix, 11, 15, 112, 115, 116, 117, 118, 123, 125, 129, 216, 219, 224, 233 fluidized bed, 5, 143, 148, 151 fluorescence, 87 fluoride, xiii, 23, 45, 215 fluorinated, 217 flushing, 60, 61, 62, 303 focusing, 243 FOG, 12 food, 12, 49, 60, 67, 199, 202, 238, 239, 240, 242, 252, 280, 300, 306, 317, 321, 329, 331, 333, 334 food industry, 60, 202, 240, 329 food production, 306 foodstuffs, 334 Forestry, 1, 279 forests, 302, 303, 304, 307 formaldehyde, 201 fossil fuels, 136, 154 fouling, 10, 219, 284 Fourier, xvi, 320, 324 fractionation, xiv, 80, 265, 273, 275, 278 France, 280, 311 freeze-dried, 182 freight, 12 fresh water, xv, 53, 280, 299, 304, 309 freshwater, 203 Freundlich isotherm, 323, 326 fructose, 281 fruits, 198, 303 FTIR, 328, 329, 330 FT-IR, 324 FT-IR, 329 fuel, 61, 69, 82, 136, 180 fumaric, 55, 57 funding, 296 fungal, xvi, 183, 185, 189, 191, 319, 320, 321, 324, 329, 331, 334 fungal metabolite, 334

Index fungi, xvi, 64, 181, 183, 184, 185, 186, 190, 192, 193, 194, 195, 241, 319, 320, 321, 324 fungicide, 142 fungus, 193, 194, 195, 212

G gas, x, xi, xvi, 7, 8, 13, 53, 61, 73, 93, 96, 98, 108, 111, 112, 113, 114, 122, 124, 125, 127, 128, 130, 132, 135, 136, 137, 138, 139, 142, 143, 147, 148, 149, 150, 152, 155, 158, 160, 161, 176, 177, 226, 236, 237, 285, 323, 333, 335 gas chromatograph, xvi, 226, 333 gaseous waste, 66 gases, xi, xiii, 7, 16, 49, 58, 59, 84, 92, 93, 95, 96, 135, 136, 137, 139, 147, 148, 149, 235, 237 gasification, 82, 108, 136 Gaussian, 117 gel, 187, 189, 190 gene, 245, 246 gene transfer, 245, 246 generation, 7, 19, 22, 108, 158, 170, 173, 211, 231, 237, 250, 252, 273, 275, 306, 320 genes, 79, 241, 242, 246 Geneva, 307, 308, 347 geochemical, 152 geothermal, 53 Germany, 45, 148, 216, 230, 232 GFP, 246 glaciers, xv, 299 glass, 64, 82, 83, 85, 106, 108, 110, 340, 341, 342 glasses, 109 global warming, 237 glucose, 54, 182, 183, 184, 185, 186, 189, 192, 193, 196, 281 glucose oxidase, 183 glycerin, 184 glycine, 182, 185, 196 glycol, 61, 114, 126, 127, 178 goals, 122, 302, 306, 307 gold, 65, 150, 280 Gore, 133 government, 231, 301, 306, 307, 309 graduate students, 296 grain, 126 grains, 63, 64 Gram-negative, 80, 241 Gram-positive, 241 granules, 64, 174 graphite, 43, 220 grasses, 154 grassland, 245, 246 gravity, 42, 87, 119, 252, 253, 275, 321

353

grazing, 238, 304 Great Britain, 296 Greece, 277 green beans, 281 greenhouse, 82 ground water, xii, 153, 173, 174, 233, 241, 302 groundwater, xii, 12, 53, 177, 197, 230, 244, 334, 344, 345, 346 groups, xvi, 2, 12, 55, 57, 59, 65, 67, 69, 70, 71, 74, 88, 91, 92, 106, 123, 127, 142, 144, 181, 185, 190, 216, 226, 236, 244, 320, 329, 330 growth, ix, xii, xv, 2, 20, 50, 51, 55, 56, 57, 59, 64, 68, 69, 73, 76, 77, 93, 101, 121, 137, 141, 143, 146, 148, 149, 151, 156, 157, 164, 175, 182, 183, 184, 191, 192, 197, 199, 202, 209, 210, 211, 239, 244, 267, 276, 299, 304, 306, 307, 321, 331 growth rate, 121, 156, 157, 164, 184, 209, 210, 276 growth temperature, 56, 57, 141 guidelines, 125, 126, 316 Gulf of Mexico, 120, 121, 125 gut, 240

H H1, 206 H2, 6, 17, 57, 137, 139, 140 habitat, 195 haloalkaliphilic, 151 halogens, 200 handling, 82, 83, 107, 237, 250, 251, 266, 273 harbour, 316 hardness, 12, 100 harm, 209 harvest, 208, 280 harvesting, 12, 240 Hawaii, 245 hazardous substance, 334, 347 hazardous substances, 347 hazardous wastes, 109 hazards, 50, 250 health, xiii, xvi, 121, 235, 240, 245, 250, 306, 308, 315, 319, 320, 333 health effects, 334 heart, 9, 230 heat, 62, 83, 101, 103, 106, 136, 193, 237, 238, 240 heating, 61, 85, 87, 89, 97, 182, 238, 240 heavy metal, x, 12, 14, 52, 62, 65, 66, 67, 71, 72, 74, 76, 77, 78, 82, 83, 84, 86, 87, 101, 102, 103, 104, 105, 106, 107, 108, 109, 112, 140, 141, 148, 151, 152, 208, 222, 239, 240, 243, 320, 330, 331 heavy metals, x, 12, 52, 62, 65, 66, 67, 71, 74, 76, 77, 78, 82, 83, 84, 86, 87, 101, 102, 103, 104,

354

Index

105, 106, 108, 109, 112, 140, 141, 148, 151, 152, 208, 222, 239, 240, 243, 320, 330, 331 height, 203, 294 hemodialysis, xv, 309, 310, 311, 312, 313, 314, 315 hemoglobin, 154 herbicide, 229, 230, 231 herbicides, 59, 201 heterotrophic, 143, 146, 155, 156, 161, 173, 202, 267 heuristic, 14 hexane, 54, 226 high temperature, 31, 65, 102, 164, 344 highlands, 280, 301, 303 hips, 27, 45 holistic approach, ix, 1, 14 homogenized, 255 homogenous, 322 horizontal gene transfer, 246 hormone, 334 hormones, 222, 233 horse, 253 hospital, 223, 245, 310, 316 hospitals, 310 host, 242 hot water, 198 hotels, xiv, 265 household, 239, 240 household waste, 239, 240 households, 59 HPLC, 340 human, xiii, xvi, 49, 121, 221, 235, 237, 240, 242, 243, 300, 319, 320, 334 human exposure, 334 humans, 202, 238, 347 humic acid, 78, 245 humic substances, 48 hybrid, 193 hybridization, 140, 147, 149 hydrate, 97, 114 hydration, 2 hydraulic fluids, 334 hydro, xi, 2, 53, 59, 61, 62, 65, 77, 78, 79, 111, 112, 113, 114, 115, 121, 124, 237 hydrocarbon, xi, 75, 111, 121 hydrocarbons, xi, 53, 59, 61, 62, 77, 79, 111, 112, 113, 114, 115, 121, 237 hydrochloric acid, 226 hydrogen, 7, 8, 17, 52, 53, 55, 57, 60, 62, 69, 70, 71, 72, 73, 74, 78, 107, 114, 137, 140, 144, 149, 150, 151, 152, 183, 186, 194, 199, 200, 201, 202, 207, 212, 219, 220, 231, 237, 285 hydrogen gas, 7, 237, 285 hydrogen peroxide, 183, 186, 194, 199, 200, 201, 202, 207, 212, 219, 231

hydrogen sulfide, 78, 107, 137, 150, 151, 152 hydrological, 300, 302, 304 hydrolysis, 5, 10, 60, 67, 69, 114, 172, 173, 236 Hydrometallurgy, 79, 150, 152 hydrophobic, 2, 3 hydroquinone, 202 hydrothermal, 56, 144 hydroxide, 7, 8, 10, 18, 31, 43, 62, 208, 226, 287 hydroxides, ix, 1, 5, 7, 13, 114 hydroxyl, xvi, 7, 8, 11, 212, 219, 220, 221, 222, 320, 329, 330 hydroxyl groups, xvi, 320, 329, 330 hypothesis, 83

I IARC, 347 identification, 59, 102, 140, 147, 242, 268, 271, 338 IEA, 237, 244 illumination, 203, 210 images, 98 immobilization, 102, 108 impact assessment, 124 implementation, 123, 163, 164, 172, 178, 198, 218, 244, 306 impurities, 11, 220, 250 in situ, xi, xiii, 7, 10, 13, 26, 111, 121, 124, 126, 216 in vitro, 347 in vivo, 347 inactivation, 233 inactive, 154, 320 incidence, 243 incineration, x, 81, 82, 107, 108, 199, 237 inclusion, x, 2 income, 306 incubation, 51, 58, 238, 239 incubation period, 238 independence, 242 India, xiv, 44, 249, 251, 259, 316 Indian, xiv, 135, 249, 250, 262 indication, 4, 15, 126, 146, 285 indicators, 57, 276 Indigenous, 280 indium, 25, 44, 297 indole, 54 Indonesia, 180 inducible enzyme, 183 induction, 186 industrial application, xi, 135, 142, 296, 320 industrial chemicals, 222, 233, 266, 347 industrial combustion, 136 industrial sectors, xiv, 265 industrial wastes, 51, 52, 65, 74, 109, 136, 201, 277

Index industrialization, xi, 135, 199 industry, ix, 2, 50, 59, 60, 61, 63, 66, 80, 112, 130, 132, 137, 147, 198, 199, 202, 203, 211, 223, 240, 272, 273, 277, 278, 309, 329 inert, xiv, 265, 269, 270, 271, 273, 276, 277, 278 infants, 154 infertile, 241 infrared, 73, 108 infrared light, 73 infrastructure, 306, 307 ingestion, 122 inhibition, 23, 76, 79, 144, 147, 157, 158, 176, 177, 186, 209, 210, 211, 230, 236, 281 inhibitors, 56, 114, 267 inhibitory, 79, 137, 138, 209, 269 inhibitory effect, 209 initiation, 104 injection, 238, 335 innovation, 11 inoculum, 167, 321 inorganic, x, 50, 51, 59, 62, 64, 65, 72, 74, 81, 82, 85, 91, 97, 98, 107, 112, 114, 144, 154, 178, 199, 219, 240 inorganic salts, 62 insect repellents, 333 insertion, 224 insight, 107, 222, 223 Inspection, 308 instability, 4, 240 instruments, 87 insulators, 295 integrated unit, 14 integration, 316, 338 interaction, 88, 115, 154, 161, 224, 322 interactions, 8, 16, 102, interface, 5 intermediaries, 201 International Agency for Research on Cancer, 347 intrinsic, 91 invertebrates, 121, 144 investment, 25, 227 ion adsorption, 4 ionic, 2, 4, 31, 173, 290 ionization, 226, 335 ions, ix, x, xvi, 1, 4, 5, 7, 8, 10, 15, 16, 17, 18, 21, 23, 25, 31, 37, 55, 56, 63, 67, 81, 95, 100, 103, 106, 126, 147, 200, 201, 211, 221, 225, 287, 289, 319, 320, 321, 322, 323, 324, 325, 326, 327, 329, 330, 331 IR, 324, 329 Iran, 300

355

iron, ix, 1, 5, 7, 9, 10, 15, 17, 20, 25, 27, 41, 43, 44, 45, 47, 55, 56, 59, 65, 67, 72, 76, 79, 92, 114, 139, 154, 201, 204, 208, 211, 284, 288, 297, 330 irradiation, 238, 281 irrigation, xiii, xiv, xv, 198, 199, 203, 208, 249, 262, 291, 309, 312, 314, 315, 317 irritation, 305 Islamic, 299 isoelectric point, 4, 5 isolation, 58, 137 isomers, 338, 339 isothermal, xvi, 319 isotherms, 322, 331 isotope, 126, 227 Israel, 300, 301 Italy, 111, 132, 170, 198, 280 IUCN, 308

J JAG, 183, 193 Japan, 87, 167, 168, 181 Japanese, 181 Jordan, 300 judge, 102

K Kenya, x, xv, 1, 2, 43, 44, 46, 47, 279, 280, 281, 282, 285, 290, 295, 296, 297 ketones, 76 killing, 56 kinetic energy, 115, 119 kinetic model, 323, 329 kinetic parameters, 158 kinetics, 37, 108, 152, 177, 212, 278, 323, 327, 331 King, 131, 179 Kobe, 184, 196 Korean, 45

L LAB, 191 labeling, 91, 99, 103 labor, 25, 42, 296, 312, 313, 314 labour, xiv, 13, 199, 249, 261 lactic acid, 68, 186, 191, 195 lactic acid bacteria, 191 Lactobacillus, 183, 185, 190, 191, 195 lactose, 60 lagoon, 75, 180 Lagrangian, 123, 127, 128

356

Index

Lagrangian approach, 123 lakes, 72, 300, 305 lamina, 33 laminar, 33 land, xiv, 12, 36, 112, 175, 236, 237, 238, 240, 241, 242, 243, 244, 245, 247, 253, 300, 301, 306, 312 landfill, 11, 44, 46, 48, 82, 109, 156, 237, 240 landfills, 30, 47, 82 Langmuir, xvi, 319, 322, 323, 326, 327, 329 lanthanoids, 65 lanthanum, 65 larvae, 121 Latin America, 306 lattice, 7, 84, 90 laundry, 11, 204 law, 41 leach, 33, 104, 287, 334 leachate, x, xv, 2, 10, 11, 17, 18, 19, 20, 22, 23, 24, 25, 28, 31, 33, 36, 40, 41, 42, 43, 44, 46, 47, 48, 237, 279, 282, 285, 286 leachates, 25, 32, 42, 48, 156, 284 leaching, x, 31, 32, 36, 37, 50, 79, 82, 83, 84, 86, 87, 101, 102, 103, 104, 106, 109, 284, 286, 287, 334 leakage, 244 leather, 61 legionella, 231 Legionella, 234 legislation, 154, 204 Libya, 301 lifestyle, 300 lifetime, 224 lift, 139, 150, 152, 176, 177 ligand, 7 lignans, 334 lignin, 12, 181, 183, 296 likelihood, 242 limitation, 14, 76 limitations, 137, 140, 170, 175, 201, 243 limiting oxygen, 160 lindane, 237 linear, 145, 209, 229, 322, 323, 339 lipid, 329 lipids, 54, 198, 236 lipophilic, 228 liquid chromatography, 226 liquid fuels, 136 liquid phase, 50, 84, 88, 93, 95, 102, 203 liquidation, 80 liquids, 176, 178 liquor, 164, 166, 193, 198, 253 Lithium, 32 loading, 19, 20, 52, 138, 140, 143, 144, 145, 146, 147, 166, 191, 251

local community, 303 localised, 45 location, 106, 110, 118, 119, 203 LOD, 226 London, 7, 44, 150, 296, 347 long period, 104, 106, 241 longevity, 10 losses, 334 Low cost, 262 low molecular weight, 113, 115 low temperatures, 22, 283 LSU, 224, 225 lubricants, 334 Luxembourg, 245 lying, 250

M machinery, xiv, 249 machines, xv, 203, 309, 310 macronutrients, 237 magmatic, 65 magnesium, 10, 30, 32, 92, 225, 234 magnetic, 192, 203 Maillard reaction, 180, 193 maintenance, xiii, 10, 11, 13, 14, 25, 152, 164, 235, 306, 312, 313, 314 Maintenance, 314 maize, 303 Malaysia, 46 malic, 53, 200 maltose, 184 management, x, xi, xiv, 2, 66, 81, 82, 111, 130, 136, 158, 164, 176, 201, 237, 245, 262, 265, 266, 271, 306, 315, 316, 317 management technology, 82 manganese, 30, 65, 67, 114, 183, 186, 296 manufacturing, 27, 45, 142, 266, 269, 271, 273, 344 manure, 239, 246, 253 marine environment, xi, 111, 112, 113, 114, 120, 121, 122, 124, 125, 126, 127, 130, 132, 161, 347 market, 8, 25, 42, 284 marketing, 280 markets, 244 Martinique, 280 mass loss, 40, 41 mass spectrometry, 226, 227 mass transfer, 140, 326 matrix, 98, 103, 104, 125, 126, 216, 222, 224, 226, 242, 334 Mauritania, 301 maximum specific growth rate, 210, 276

Index MDA, vii, xii, 179, 180, 181, 182, 183, 184, 185, 186, 188, 189, 190, 191, 192 measurement, 112, 322 measures, 117, 119, 125, 126, 136 meat, 60, 277, 280 media, xi, 4, 54, 57, 59, 83, 135, 182, 216, 231, 281 Mediterranean, xi, xii, 111, 113, 120, 121, 124, 127, 197, 198, 199, 245 Mediterranean countries, 198, 199 melting, 93, 94, 95, 98 membranes, 113, 310, 311 Merck, 339 mercury, 65, 76, 114, 237 metabolic, 52, 75, 140, 145, 242, 273 metabolism, 51, 55, 67, 71, 75, 147, 152, 240, 241 metabolites, 77, 229, 230, 231, 334 metal content, 72, 239 metal hydroxides, 5 metal ions, 5, 8, 15, 16, 21, 95, 106, 147, 201, 289, 324, 331 metal oxide, 82, 85, 223, 287 metal oxides, 82, 287 metallurgy, 66, 67 metals, x, 12, 52, 62, 65, 66, 67, 71, 72, 74, 76, 77, 78, 79, 82, 83, 84, 86, 87, 101, 102, 103, 104, 105, 106, 107, 108, 109, 110, 112, 114, 121, 140, 141, 148, 151, 152, 208, 222, 237, 239, 240, 243, 244, 320, 322, 330, 331 methane, xiii, 69, 82, 141, 170, 172, 191, 235, 236, 237, 240 methanogenesis, 63, 70, 71, 79, 147, 212 methanol, 54, 157, 164, 173, 226, 335 methemoglobinemia, 154 methylation, 76 methylene, 331 Mexico, 120, 121, 125 microaerophilic, 138, 139 microalgae, xiii, 198, 202 microbes, 180, 192 microbial, xi, xii, xiii, xiv, 75, 78, 79, 115, 135, 137, 143, 147, 148, 149, 150, 154, 177, 179, 181, 182, 185, 186, 190, 191, 192, 197, 204, 235, 236, 238, 239, 241, 242, 243, 244, 245, 246, 247, 266, 270, 271, 275, 278, 304, 311, 321 microbial agents, 238 microbial communities, xiii, 79, 235, 241, 242, 243, 245, 246 microbial community, xiv, 115, 236, 239, 241, 245, microbiota, 239 microcosms, 57, 58, 77, 80 microflora, 55, 199 micrograms, 334 micronutrients, 237

357

microorganism, 51, 52, 53, 57, 59, 64, 67, 69, 70, 182, 183, 206, 267 micro-organisms, 115, 245, 246 microscope, 87, 92 microstructures, 92,93, 98 Middle East, 300, 301, 317 migration, 21, 23 military, 300, 301 milk, 50, 60, 61, 62 Millennium, 302, 308 milligrams, 37 mineral oils, 62 mineral water, 339, 341, 342, 343, 344, 346 mineralization, 52, 74, 76, 141, 201, 221, 239, 242 mineralized, 220 minerals, 30, 79, 90, 98, 102, 137, 180, 291 Minerals Management Service, 132 mines, 62 mining, xvi, 49, 82, 319, 320 Minnesota, 79 misleading, 267, 269 missions, 136, 168, 173 mixing, ix, xi, 7, 15, 33, 86, 111, 112, 113, 115, 116, 118, 119, 121, 124, 127, 129, 130, 131, 169, 209, 218, 228, 231, 236, 251 MMS, 132 mobility, 3, 57, 65, 79, 104, 106 MOD, 133 model system, 193, 196 modeling, 278 models, 120, 122, 123, 124, 126, 131, 209, 266, 275, 315, 323, 326 modulus, 85 moisture, 199, 230, 237, 245, 302 moisture content, 199, 245 molar ratio, 160 molasses, xii, 48, 63, 179, 180, 181, 182, 185, 186, 189, 190, 191, 192, 193, 194, 195, 196 mold, 181 mole, 293 molecular weight, 113, 114, 115, 180, 181, 184, 186 molecular weight distribution, 181, 186 molecules, 10, 140, 322 molybdenum, 65 momentum, 15, 115, 118, 119, 120, 128 money, 250 monolayer, xvi, 319, 329 monomeric, 70 Moon, 108 Morocco, 300, 309, 310, 312, 316 morphological, 83, 84, 88, 96, 99 morphology, 92, 98, 126, 145 mortality, 154, 175

358

Index

motion, 2, 4, 13, 119 motors, 261 mountains, 304 movement, 22, 341 multidisciplinary, xi, 111, 113, 126, 127, 130 multidrug resistance, 245 multiplication, 57, 58 municipal sewage, xiii, 82, 221, 235, 244, 311 municipal solid waste, 82, 109 mycelium, 188

N Na+, 4, 31 Na2SO4, 23, 24, 63 NaCl, x, 2, 23, 24, 25, 62, 321, 336, 339, 341, 346 Namibia, 300 naphthalene, 58, 121, 277 National Bureau of Standards, 94 native plant, 245 natural, x, xi, xiii, 5, 9, 12, 22, 44, 46, 47, 49, 50, 51, 53, 57, 58, 59, 65, 66, 71, 72, 73, 74, 75, 83, 93, 111, 113, 121, 122, 123, 141, 152, 154, 181, 198, 204, 241, 250, 251, 273, 277, 300, 306, 307, 310, 320, 321, 330 natural environment, 49, 50, 51, 53, 57, 58, 65, 66, 71, 72, 73, 74 natural gas, xi, 53, 111, 113, 152 natural resources, 83 Nb, 216, 218 NCA, 46 Near East, 277, 278 neck, 93 neem, 331 Nepal, 316 Netherlands, 87, 121, 147, 149, 158, 163, 164, 167, 168, 308 network, 88, 92, 95, 100, 106 neutralization, 7, 47, 50, 66, 203, 207, 208 New Frontier, 77 New Jersey, 245 New York, 45, 46, 75, 80, 131, 132, 133, 151, 152, 165, 196, 234, 262, 276, 300, 308 Ni, 52, 85, 331 nickel, 65, 66, 67, 109, 114, 330 Nielsen, 53, 78 Niger, 300, 301 Nile, 300, 301, 304 niobium, 65 nitrate, 11, 23, 78, 143, 149, 154, 155, 156, 157, 161, 162, 163, 171, 173, 174, 175, 177, 241, 297 nitrates, 57 nitric oxide, 150, 177

nitrification, xi, 50, 153, 155, 156, 157, 158, 159, 160, 162, 164, 166, 167, 168, 169, 170, 175, 176, 177, 178, 241, 243 nitrifying bacteria, 156, 162 Nitrite, 154, 158, 160, 164 nitrogen, xi, xii, 49, 50, 61, 75, 153, 154, 155, 156, 158, 159, 160, 161, 162, 165, 168, 170, 173, 174, 175, 176, 177, 184, 211, 226, 237, 239, 243, 250, 305, 306, 308, 311, 312, 313, 320 nitrogen compounds, xii, 49, 153, 154, 161, 237 nitrogen fixation, 75, 154 nitrogen gas, 155, 158, 160, 161, 176 nitrogen oxides, 161 nitrosamines, 154 nitrous oxide, 177 non toxic, 130 non-enzymatic, 180 normal, xii, 12, 61, 157, 179, 180, 190, 219, 238, 285, 289 norms, 51 North Africa, 46, 300, 317 North America, 300 Norway, 121 NTU, 22 nucleic acid, 147 numerical tool, 123 nutrient, xi, xiii, 115, 144, 146, 153, 202, 235, 236, 246, 307, 321 nutrient cycling, xiii, 235 nutrients, xiii, 126, 142, 198, 211, 237, 238, 239, 242, 243, 250, 303, 320 nutrition, 202, 212

O observations, xi, 92, 98, 111, 112, 113, 122, 125, 126, 130 octane, 54 odors, 140, 211, 238 offshore, x, xi, 111, 112, 113, 114, 115, 120, 121, 122, 124, 126, 127, 130, 132 offshore oil, x, 111, 112, 127, 130, 132 oil, xi, xii, xiii, 11, 44, 50, 53, 54, 57, 61, 62, 73, 77, 78, 79, 80, 107, 108, 111, 112, 113, 114, 122, 124, 125, 127, 130, 132, 136, 151, 197, 198, 199, 202, 203, 204, 205, 206, 208, 209, 211, 212, 215, 228, 229, 239, 244, 245, 300 oil production, 198, 199 oil refineries, 50, 62 oil refining, 136 oil shale, 11 oils, 62

Index olive, ix, xii, 197, 198, 199, 202, 203, 204, 205, 206, 208, 209, 211, 212, 213, 330, 331 olive oil, xii, 197, 198, 199, 202, 204, 205, 206, 212 olives, xii, 197, 198, 202, 203, 204, 208, 209 Oman, 300 online, 151, 308 Operators, 125 optical, 73, 78, 182 optical density, 182 optimization, 14, 22, 147, 227, 231 oral, 122 ores, 66, 72, 150 organelles, 188 organic C, 241 organic compounds, 10, 52, 53, 54, 55, 56, 57, 58, 59, 60, 61, 63, 67, 68, 69, 70, 72, 73, 74, 113, 140, 142, 144, 200, 211, 213, 237, 262, 267, 334 organism, 139, 145 organochlorinated, 201 orientation, 120 oscillation, 52 osmosis, 12, 173, 310, 311, 313, 314, 315, 320 osmotic, 9 osmotic pressure, 9 ovulation, 230 oxalic, 201 oxalic acid, 201 oxidants, 221, 222 oxidation rate, 145, 158 oxidative, x, 19, 82, 84, 104, 106, 201, 220 oxide, x, 10, 13, 30, 43, 53, 82, 83, 85, 95, 97, 98, 99, 102, 150, 177, 208, 217, 223, 228, 231, 330 oxide electrodes, 10 oxides, x, 13, 30, 81, 82, 83, 84, 85, 88, 95, 96, 100, 102, 161, 218, 287, 289 Oxygen, 47, 145, 151, 160, 178, 219, 262 oxygenation, 59, 67, 150 ozonation, 222, 234, 277 ozone, xiii, 49, 109, 154, 216, 219, 222, 224, 225 Ozone, 219

P Pacific, 131, 132 packaging, 333 PAHs, 113 paints, 66, 320 Palestine, 300, 301 palletized, 109 paraffins, 201 parameter, 15, 17, 18, 33, 116, 119, 120, 141, 203, 225, 226, 266, 267, 268, 269, 290 parasites, 246

359

Paris, 334, 347 Parliament, 233, 347 particle density, 83, 87, 88, 89, 95, 96 particles, xiii, 2, 4, 5, 7, 11, 13, 31, 33, 46, 64, 66, 88, 93, 100, 103, 113, 123, 142, 144, 174, 215, 217, 252, 290 particulate matter, 108, 121 partition, 114, 198, 255 passivation, 10, 13, 217 passive, 10, 49, 112 pasteurization, 240 pasture, 307 pathogenic, xiii, 52, 71, 83, 235, 236, 237, 240, 242, 243, 245, 246 pathogenic agents, 83 pathogens, 82, 175, 238, 240, 241, 244, 250, 306 patients, 310 Pb, viii, x, xvi, 21, 82, 84, 86, 101, 102, 103, 104, 106, 290, 319, 320, 321, 322, 323, 324, 325, 326, 327, 328, 329, 330, 331 PCBs, 201, 334 PCT, 36, 46, 233, 296 peat, 262, 331 pectin, 198, 281 per capita, 300 percolation, 253, 262 periodic, 14 permeability, 113 permeable membrane, 121 permit, 33 permittivity, 3 peroxide, 183, 186, 194, 199, 200, 201, 202, 206, 207, 212, 219, 231 pesticide, 66, 234, 347 pesticides, 59, 66, 202, 209, 211, 305, 340 PET, 342, 344 petrochemical, 43, 50, 53, 54, 57, 58, 61, 74, 170 petroleum, 77, 78, 80, 280 Petroleum, 61 petroleum products, 80 Petrology, 49 pH values, 9, 17, 141, 324 pharmaceutical, 69, 226, 233, 234 pharmaceuticals, xiii, 215, 222, 225 PHB, 202 phenol, 54, 58, 61, 62, 75, 78, 202, 203 phenolic, xii, 193, 197, 198, 204, 208, 209, 211, 212 phenolic acid, 193 phenolic compounds, xii, 197, 198, 204, 208, 209, 211, 212 Philippines, 180 philosophy, xiv, 265, 266, 271

360

Index

phosphate, x, xv, 2, 12, 36, 39, 43, 44, 47, 241, 279, 280, 282, 284, 285, 286, 287, 289, 292, 296, 297 Phosphate, xv, 24, 279, 284 Phosphogypsum, 80 phosphorous, 305 phosphorus, 50, 107, 108, 172, 250, 311, 313 phosphorylation, 55 photobioreactors, 203 photoionization, 347 photolysis, 113 photomicrographs, 93 photosynthetic, 12, 78, 137, 143, 150, 202 phototrophic, 77, 142, 144 phylum, 158 physical factors, 14 physical properties, 238, 244, 246 physicochemical, 114, 136, 164, 251, 255 physiological, 59, 246 physiology, 53, 145, 308 phytoplankton, 121, 304 pig, 239 pigments, 195 pipelines, 61 planar, 88 planning, xi, 111 plants, xi, xii, xiii, 30, 49, 51, 59, 60, 61, 62, 66, 73, 74, 75, 86, 136, 144, 153, 154, 157, 167, 194, 197, 222, 233, 235, 236, 239, 240, 244, 245, 246, 266, 267, 271, 311, 317, 321 plastic, xiii, 64, 66, 215, 217, 284 plasticizer, 334 plastics, 61, 69, 73 platforms, xi, 111, 113, 114, 115, 120, 121, 122, 124, 125, 126, 127, 128, 130 play, xvi, 63, 67, 95, 120, 125, 141, 147, 304, 320 Pleurotus ostreatus, 183 ploughing, 240 Poland, 49 polarity, 10, 11, 14, 115, 219, 224, 229, 341 politics, 300 pollutants, ix, xi, xiii, 1, 2, 7, 8, 12, 19, 20, 22, 49, 50, 51, 57, 58, 59, 61, 62, 63, 67, 69, 74, 83, 107, 108, 116, 117, 123, 130, 135, 200, 201, 215, 216, 237, 239, 243, 250, 320, 334, 339, 341, 342, 346 pollution, x, xi, xv, xvi, 2, 12, 13, 42, 49, 50, 51, 64, 73, 82, 83, 107, 111, 117, 136, 149, 154, 200, 231, 237, 241, 262, 267, 282, 299, 300, 304, 305, 306, 307, 319, 333, 335, 339, 340, 346 polyamide, 278 polycyclic aromatic hydrocarbon, 78, 124 Polyelectrolyte, 28, 30 polyethylene, 178 polymer, 66, 145, 334, 344, 346

polymer synthesis, 344 polymerization, 12 polymers, 196 polyphenolic compounds, 199 polyphenols, 199, 206, 212 polysaccharides, 188, 329 polyurethane, 190, 193 polyurethane foam, 190, 193 polyvinyl chloride, 334 pomace, 198, 199 pond, xii, 179, 180, 181 pools, 208 poor, xii, 146, 179, 184, 239, 303, 304, 306, 310 population, ix, xv, 2, 67, 146, 280, 299, 300, 301, 304, 306, 307, 315 population growth, ix, xv, 2, 299, 304, 306, 307 pore, 93, 95, 203, 240 pores, x, 81, 84, 88, 92, 93, 95, 98, 100, 104, 106 porosity, 83, 87, 88, 89, 95, 96, 98, 240, 243 porous, 92, 93, 100, 102, 106, 194 ports, 119, 120 potassium, 30, 37, 114 potato, 27, 45, 188, 321 potatoes, 303 poultry, 276 poverty, 303 powder, 30, 87, 92, 102, 182, 331 powders, 324 power, xi, xv, 9, 12, 17, 20, 22, 23, 24, 26, 28, 30, 36, 41, 42, 62, 136, 169, 221, 223, 228, 233, 237, 253, 258, 279, 282, 284, 287, 288, 289, 291, 293, 295, 296, 300, 317, 323 power plants, xi, 30, 62, 136 precipitation, 5, 53, 59, 66, 67, 71, 114, 208, 219, 224, 302, 303, 320, 324 preconditioning, 12 prediction, ix, 1, 14 preference, 9, 115 press, 44, 84, 87, 94, 151, 199, 246, 297, 308 pressure, ix, 2, 9, 61, 65, 136, 174, 246, 300, 303, 304, 307 prevention, 125 prices, 25, 201, 313, 314 printing, 66 probability, 92, 98 process control, 330 producers, xii, xiv, 30, 197, 199, 216, 217, 236, 303 production costs, xv, 309 productivity, 137 profit, 55 program, 226, 311, 316, 317, 321 prokaryotes, 80 prokaryotic, 140

Index proliferation, 156, 242 propylene, 61 protection, 50, 74, 136, 307, 347 protein, 12, 56, 60, 154, 329 proteinase, 241 proteins, 54, 55, 60, 142, 180, 236, 242, 306 Proteins, 61 protocol, 114, 130, 266 protocols, 268, 271 protons, 161, 324 protozoa, 237 pseudo, xvi, 14, 18, 320, 323, 327, 328, 329 Pseudomonas, 145, 185, 190, 193, 194, 246 public, 82, 123, 126, 306, 316 public administration, 126 public health, 306 pulp, ix, 2, 8, 12, 17, 20, 23, 40, 41, 43, 44, 47, 61, 69, 136, 198, 276, 277, 281, 282, 296, 297 pulp mill, 296 pumping, 169, 175, 232, 252, 257, 296 pumps, 223, 261 purification, xiii, 44, 109, 198, 207, 233 PVC, 333, 334, 344 pyrolysis, 82, 107, 108 pyruvic, 53, 55, 57

Q Qatar, 300 quality of life, 250 quartz, 91, 92, 98, 99, 102, 103 quinones, 201

R race, 237 radiation, 73, 87 rail, 229 rain, 59, 136, 302 rain forest, 302 rainfall, 230, 301, 302 rainwater, 79 random, xv, 4, 279 range, ix, x, 2, 8, 10, 12, 14, 17, 18, 25, 49, 50, 51, 56, 70, 74, 81, 88, 90, 95, 96, 122, 123, 125, 136, 138, 139, 140, 141, 144, 146, 173, 207, 210, 217, 220, 222, 227, 236, 245, 252, 273, 285, 326, 338, 339, 342 rangeland, 246 rare earth, 50, 65 rare earth elements, 65

361

raw material, xi, xii, 61, 83, 84, 85, 87, 88, 89, 91, 92, 93, 95, 102, 135, 179, 180, 191, 192 raw materials, xii, 83, 84, 85, 87, 88, 89, 91, 92, 93, 95, 102, 179 reaction order, 33 reaction rate, 13, 19, 21, 22, 37, 40, 328 reaction time, 21, 296 reactivity, 100, 109, 284, 289 reagent, 199, 200, 201, 202, 212, 219, 276, 322 reagents, xvi, 25, 333, 335, 339, 341, 346 receptor sites, 8 receptors, 334 reclamation, 142, 148, 240, 317 Reclamation, 311 reconditioning, 12 recovery, xvi, 13, 65, 66, 72, 77, 83, 137, 146, 150, 151, 172, 212, 244, 262, 278, 310, 316, 333, 341 rectification, 61 recycling, xiv, xv, 12, 82, 83, 156, 236, 237, 238, 244, 280, 296, 303, 309, 310, 313, 314, 315 redox, 17, 55, 57, 58, 59, 145, 150, 219, 225, 239 Redox, 226 reducing sugars, 281 REE, 65 refineries, xi, 50, 62, 66, 136, 200 refining, 53, 61, 62, 78, 80, 109, 136, 320 refractory, 150 refrigeration, 344 regeneration, 82, 137 regional, 122, 124, 309 regression, 328 regulation, 67, 130, 222 regulations, 8, 82, 125, 208, 244, 281 rejection, 114, 178 relationship, 3, 9, 18, 31, 41, 69, 70, 75, 104, 107, 119, 202, 207, 241 relationships, 15, 52, 59, 69 reliability, 123 remediation, 43, 193 renal, 310 renal failure, 310 renal function, 310 renewable energy, 136, 237, 244 reproduction, 154 reserves, 112 reservoirs, x, 53, 59, 78, 111, 113, 114, 246 residential, 136, 310 residuals, 107, 108 residues, 82, 107, 220, 226, 239, 240, 246 resins, 61, 62, 333, 334 resistance, 10, 25, 27, 77, 101, 140, 222, 240, 242, 245, 246, 326 resolution, 130

362

Index

resources, ix, x, xi, xv, 2, 81, 82, 111, 123, 177, 299, 300, 301, 302, 304, 305, 307 respiration, 52, 53, 56, 76, 163, 239 respiratory, 140, 239 restaurant, 11 restaurants, 239 retention, 27, 139, 150, 158, 164, 168, 173, 239, 251, 252, 254, 257, 259, 321, 338, 339 Reynolds, 15, 347 Reynolds number, 15 rice, 304 rings, 142 risk, 130, 170, 230, 233, 236, 237, 238, 240, 241, 242, 243 risks, xiv, 83, 112, 124, 125, 236, 240, 243, 244, 246 river systems, 303 rivers, 199, 233, 300, 302, 303, 304, 305, 307 rods, 5, 151 Rome, 111, 316 room temperature, 87, 108, 146, 178, 202, 321 room-temperature, 87 Royal Society, 47 rubber, 58, 61 runoff, 64, 127, 236, 238 rural, 250, 262, 304, 305, 306, 308 rural areas, 305 rural communities, 304, 306 rust, 211

S safe drinking water, 306 safety, 84 saline, 11, 18, 72, 317, 321 salinity, ix, 1, 126, 128, 177, 178 salt, 23, 148, 184, 200, 207, 225 salts, 23, 61, 62, 83, 126, 177, 184, 201, 204, 212 saltwater, 120 sample, xvi, 4, 5, 87, 221, 224, 225, 228, 232, 256, 267, 284, 333, 345, 346 sampling, xv, 20, 40, 41, 112, 121, 125, 126, 130, 225, 255, 279, 345 sand, 11, 14, 18, 64, 82, 109, 208, 226, 330 sanitation, 59, 250, 306, 308, 315 saturation, 15, 160, 324, 329 Saudi Arabia, 1, 279, 300 savings, 164, 211, 310, 314, 315 sawdust, 262, 330, 331 SBR, 176, 178 scaling, 294 Scanning electron, 87 scarcity, 299, 300, 307, 310, 315, 316 sea level, 231

sea urchin, 121, 122 search, 51, 59, 82, 307 searching, 83, 301 seawater, xvi, 7, 113, 114, 115, 120, 121, 126, 127, 309, 312, 314, 315 security, 306 sediment, 115, 122, 125, 126, 207 sedimentation, 5, 15, 61, 65, 83, 107, 113, 114, 204, 208, 252, 253 sediments, 53, 115, 122, 127, 151, 207, 211, 334 seed, 46, 253, 267, 321 seeds, 44, 45 seeps, 223 selecting, 88, 123, 339 selectivity, 320 SEM, 87, 92, 98 semiarid, 245, 246 semi-arid, 245 semi-arid, 246 semi-arid, 246 semi-arid, 315 semiconductor, ix, 2, 11 Senegal, 301 sensitivity, 67, 139 sensors, 223, 224, 225 separation, 16, 18, 25, 46, 61, 64, 114, 115, 203, 204, 207, 211, 253, 315, 317 sequencing, 241 series, 7, 26, 47, 50, 121, 154, 207, 222, 223, 224, 225, 236, 294, 321 settlers, 168 sewage, xiii, xv, 7, 12, 49, 50, 51, 53, 57, 58, 59, 60, 61, 62, 63, 64, 66, 67, 69, 70, 71, 72, 73, 74, 75, 77, 82, 107, 108, 109, 149, 176, 206, 221, 232, 233, 235, 236, 237, 239, 240, 241, 242, 243, 244, 245, 246, 265, 266, 268, 273, 278, 304, 306, 311 Shanghai, 86 shape, 11, 57, 88, 91 shaping, 265 sharing, 251, 252, 254, 256, 257 shear, 2, 3, 120, 128, 129 shelter, 304 shock, 252 short period, 204 shortage, 300, 303, 315, 316 short-term, 67, 77, 84 sign, 2, 119, 220 silica, x, 12, 81, 92, 106, 109, 330 silicate, x, 31, 81, 85, 87, 88, 90, 91, 97, 100, 102, 106, 109 silicates, 30, 98, 100, 102, 103, 104 silver, 65 similarity, 15

Index simulations, 123, 130, 313, 315 sintering, x, 82, 83, 84, 85, 88, 89, 91, 92, 95, 98, 100, 101, 102, 103, 104, 106, 109, 136 SiO2, x, 39, 81, 83, 85, 86, 88, 89, 90, 91, 92, 93, 94, 97, 102, 106, 109, 110 SIR, 227 sites, 8, 50, 79, 237, 240, 255, 324, 326 skeleton, 88 skills, 123 skin, 93, 154, 305 slag, 64, 82 smelters, 62, 66 smelting, 320 SO2, xi, 96, 135, 136, 137, 138, 139, 147, 148, 151 sodium, 31, 57, 62, 85, 87, 92, 114, 139, 225, 226 sodium hydroxide, 62, 225, 226 software, 87, 123 soil, xi, xiii, xiv, 36, 50, 54, 66, 80, 82, 109, 135, 142, 154, 185, 199, 201, 203, 230, 234, 235, 236, 237, 238, 239, 240, 241, 242, 243, 244, 245, 246, 247, 281, 303, 316 soil erosion, 303 soils, 53, 54, 57, 80, 109, 199, 241, 242, 245, 246 solar, 141 solid phase, 64, 203, 226, 324, 326 solid waste, 50, 51, 52, 63, 66, 71, 74, 82, 109, 198, 199, 277, 304 solidification, 101, 102, 104, 106, 107 sols, 2 solubility, 16, 39, 65, 104, 106, 110, 115, 288 solvent, 50, 137, 320, 346 solvents, 50, 201, 226 Somalia, 300 sorption, xvi, 66, 108, 234, 320, 322, 323, 324, 326, 330, 331 sorption isotherms, 322 sorption kinetics, 323 South Africa, 282 sovereignty, 301 soy, 311 Spain, 153, 197, 198, 202, 204, 205, 208, 212, 213, 309, 333, 341 spatial, 59, 113, 125, 126 speciation, 16, 83, 122 species, xiii, 5, 7, 16, 33, 54, 56, 57, 68, 69, 72, 107, 121, 122, 126, 145, 146, 152, 175, 183, 200, 235, 238, 241, 245, 250, 280, 313 specific adsorption, 5 specific gravity, 87 specific surface, 7 spectrophotometry, 311 spectroscopy, xvi, 333 spectrum, ix, xiv, 57, 87, 265, 267, 329, 337, 338

363

speed, 51, 64, 83, 87, 94, 126, 128, 192, 200, 201, 210, 251, 259 sperm, 121 spheres, 5 spills, 251 spin, xiii, 215, 216 spindle, 140 spore, 321 springs, 53 SPSS, xv, 279, 283 Sri Lanka, 47 SRT, 321 stability, xiii, 2, 4, 5, 10, 11, 16, 91, 92, 95, 101, 103, 106, 108, 128, 129, 216, 217, 220, 224, 226, 239 stabilization, x, 67, 82, 84, 86, 101, 104, 107, 109, 238, 307, 308 stages, 61, 63, 64, 65, 68, 69, 71, 74, 90, 147, 188, 198, 281 standard deviation, 203 standards, xiv, xv, 136, 249, 265, 266, 271, 280, 281, 287, 306, 311, 312, 313, 315, 337 Standards, 94, 290, 296, 336 statutory, 222 steady state, 14 steel, 10, 53 sterile, 311, 321 sterols, 334 stock, 49, 322, 335, 341 stoichiometry, 69 storage, 49, 199, 208, 240, 321 storms, 154 strain, 54, 75, 77, 79, 145, 151, 181, 182, 183, 184, 185, 187, 191, 194, 195 strains, xii, 54, 58, 76, 141, 179, 182, 183, 185, 189, 190, 191, 242 strategies, 14, 159, 160, 170, 173, 243, 307, 347 stratification, 112, 119, 120, 127, 128, 130 streams, xiv, 12, 66, 76, 142, 147, 149, 152, 208, 253, 265, 266, 268, 269, 271, 302, 304, 330 strength, x, 2, 4, 31, 81, 83, 84, 87, 88, 91, 94, 95, 96, 98, 100, 102, 106, 176, 193 Streptomyces, 185 stress, 148, 150, 300 stroke, 231 students, 296 subgroups, 269 Sub-Saharan Africa, 300, 306 substances, x, 13, 48, 49, 51, 57, 81, 82, 91, 96, 97, 115, 116, 117, 126, 127, 132, 144, 180, 195, 198, 205, 219, 221, 222, 227, 228, 234, 269, 313, 334, 345, 347 substitutes, 83 substitution, 100, 106, 271

364

Index

substrates, 23, 55, 69, 140, 144, 180, 236 sucrose, 184, 281 Sudan, 301, 307 suffering, 310 sugar, 60, 63, 180, 182, 183, 185, 192, 194, 196, 280 sugar beet, 194 sugar cane, 63, 180, 185 sugarcane, 194, 251 sugars, 198, 236, 281 sulfate, xi, 23, 75, 76, 77, 78, 79, 80, 84, 135, 136, 137, 138, 139, 140, 141, 142, 143, 145, 146, 147, 148, 149, 150, 151, 152, 161, 162, 170, 171, 173, 175, 311, 313 sulfites, 147 sulfur, xi, 75, 76, 77, 135, 136, 137, 138, 139, 142, 143, 144, 145, 146, 147, 149, 150, 151, 152, 161, 162, 173, 174, 175, 176, 237, 241 sulfur dioxide, 147, 149, 150 sulfuric acid, 10, 136 sulphate, ix, 52, 53, 55, 56, 57, 58, 59, 63, 67, 68, 69, 72, 76, 77, 78, 79, 80, 138, 139, 140, 141, 145, 146, 148, 149, 150, 151, 226 sulphur, 51, 52, 53, 55, 56, 57, 59, 62, 63, 65, 69, 70, 71, 72, 73, 74, 75, 77, 79, 114, 139, 144, 145, 146, 147, 148, 149, 150, 151, 177, 243 summer, xii, 120, 127, 128, 129, 130, 197, 199 sunlight, 154 supercritical, 110 supernatant, 87, 163, 167, 172, 178 supervision, 244 supply, 50, 51, 66, 72, 92, 106, 136, 137, 160, 163, 217, 223, 228, 231, 233, 237, 239, 250, 284, 302, 304, 306, 308, 315, 344 supply chain, 237 surface area, xv, 7, 15, 108, 251, 252, 279, 284, 294, 295, 296, 326 surface component, 188 surface properties, 142 surface tension, 15 surface water, xvi, 12, 46, 51, 82, 127, 222, 233, 238, 306, 333, 334, 335 surrogates, 226 survival, 240, 245, 246, 250 suspensions, 46, 50, 62, 112 sustainability, xiii, 202, 235, 307, 315 sustainable development, x, 82, 83 sweat, 204 Sweden, 235, 240 switching, 136 Switzerland, 46, 167, 216 syndrome, 154 synthesis, 216, 331, 344 Syria, 300

T Taiwan, 193 tandem mass spectrometry, 226 tanks, 18, 50, 61, 64, 156, 163, 251, 281, 295 tannin, 14, 273, 277, 330 tannins, 198, 277 tantalum, 65 Tanzania, 284 tar, 11 taste, 83 taxonomic, 71 tea, ix, 2, 11, 12, 46 technical assistance, 330 Teflon, 339 temporal, 113, 115, 125, 126 tension, 15, 114, 304 testimony, 300 textile, ix, 2, 25, 43, 45, 61, 66, 69, 199, 269, 276, 277, 278, 317 textile industry, 61, 278 textiles, 320 TGA, 84 Thailand, 179, 180, 185, 193, 246 thallium, 65 therapy, 315 thermal analysis, 83, 84, 88, 96 thermal properties, 96, 97 thermal resistance, 11 thermal treatment, 82, 101, 102, 104, 136 thermodynamic stability, 101 thermodynamics, 331 thermogravimetric, 87 thermogravimetry, 108 thermophiles, 141 thin film, 313 Third World, 30 threat, xii, 197, 300, 309 threatened, 300 threats, xv, 299, 302 three-dimensional, 124, 128 three-dimensional model, 124 threshold, 231 thresholds, 222 timing, 240 tin, 10, 65 tissue, 121 titanium, 11, 136, 217, 224, 228 Titanium, 32 titanium dioxide, 136 titration, 225 TOC, 70, 267, 268 Tokyo, 234

Index toluene, 54, 58, 75, 76, 80, 121, 126, 146, 149, 268 tomato, 303 total energy, 136 total organic carbon, 107, 243, 267 total organic carbon (TOC), 267 toxic, xi, xvi, 7, 11, 51, 52, 56, 60, 66, 67, 69, 71, 74, 79, 104, 111, 112, 113, 120, 122, 126, 130, 141, 142, 147, 154, 173, 175, 201, 219, 239, 240, 243, 244, 269, 319, 320, 334, 347 toxic effect, xi, 111, 112, 120, 122, 126, 130, 154 toxic metals, 67, 104 toxic products, 201 toxicity, xii, 12, 67, 84, 87, 114, 120, 121, 126, 146, 152, 170, 175, 197, 209, 271, 276, 277, 278, 281 toxins, 305 toys, 334 trace elements, 114, 320 tracers, 121, 126, 127 tracking, 304 trade, 42, 216 trademarks, 202 trade-off, 42 trading, 280 trajectory, 120, 125, 128 transfer, 11, 13, 103, 140, 154, 201, 242, 243, 245, 246, 251, 252, 253, 267, 326 transformation, 50, 65, 66, 71, 75, 91, 92, 96, 97, 102, 103, 114, 344, 346 transformations, 68, 119, 154, 241 transition, 90, 201 transmission, 238 transparent, 211 transport, 5, 55, 56, 61, 66, 112, 114, 116, 117, 122, 123, 124, 126, 272, 277 transport processes, 123 transportation, xiv, 42, 65, 249, 261, 280, 333 traps, 302 travel time, 12 treatment methods, xiii, 12, 50, 59, 215 trial, 239 tribal, 307 tribes, 54, 55, 71 trichloroethylene, 201, 212 tropical areas, 180 tumours, 154 Tunisia, 300 turbulence, 16, 114, 115, 116, 120, 124 turbulent, 115, 118 Turbulent, 131 turbulent mixing, 118 Turkey, 265, 280, 319 Turku, 80

365

U Uganda, xv, 299, 301, 303, 304, 305, 306, 307, 308 ultraviolet, 281 ultraviolet irradiation, 281 UNDP, 300, 301 UNEP, 281, 297, 304, 305 UNESCO, 283, 297 UNICEF, 306, 308 uniform, 252 United Arab Emirates, 300 United Kingdom, 203, 297, 334, 347 United Nations, 311, 312, 316 United States, 7, 312 universities, 216 uranium, 79 urban areas, 303, 304, 306 urban centres, 306 urbanized, 250 urine, 236, 240 USEPA, 87, 121, 133, 237, 334

V valence, 4, 92, 106 validation, 231 values, 4, 9, 15, 17, 60, 89, 94, 104, 119, 128, 129, 141, 158, 180, 201, 203, 204, 206, 207, 208, 210, 211, 232, 240, 243, 266, 267, 286, 287, 300, 315, 321, 324 van der Waals, 5, 31 van der Waals forces, 31 vanadium, 65 vapor, 31, 216, 218 variability, 16, 127, 213, 322 variables, 25, 33, 41 variance, 283 variation, 18, 88, 92, 106, 147, 207, 210, 287, 344 vegetables, 303 vegetation, 198, 199, 301, 302, 303 velocity, 114, 115, 118, 119, 120, 144, 210, 252, 290 versatility, 136 vessels, 58, 60 Victoria, 303, 304, 305, 308 village, 250 vinasse, 48 vinyl chloride, 334 violence, 304 virulence, 240 viruses, 9, 237 viscosity, 3, 93, 95, 98 vitamins, 60

366

Index

vitrification, 108, 110 voids, 87, 88, 93, 95 volatilization, x, 81, 97, 112

W war, 300 Warsaw, 49 waste disposal, ix, 219, 304 waste incinerator, 109 waste management, 66, 82, 237, 306 waste products, 239 waste treatment, 78, 139, 150 wastes, ix, 2, 44, 51, 63, 65, 66, 74, 82, 109, 110, 140, 199, 212, 233, 236, 244, 246, 305 wastewater treatment, ix, x, xii, xiv, 2, 7, 11, 12, 27, 37, 44, 81, 86, 145, 149, 151, 153, 157, 174, 179, 180, 185, 191, 192, 197, 236, 244, 249, 250, 251, 254, 255, 256, 258, 259, 260, 261, 262, 276, 289, 306, 307, 308, 310, 311, 314, 315, 317, 320 water absorption, 87, 88, 93, 95, 98 water policy, 233, 334, 347 water quality, xv, 74, 175, 231, 299, 303, 307, 311, 316, 347 water resources, ix, xv, 2, 299, 301, 302, 305, 307 water supplies, 43 water table, 304 water vapor, 31 water-soluble, 114, 122 wealth, 42 wear, 220 weathering, 65, 114 web, 245 weight gain, 230 weight loss, 89, 91, 96, 97 weight ratio, 95 welding, 217, 295 wells, 12, 282, 302, 304 West Indies, 280 wetlands, xv, 299, 303, 304, 307 wheat, 242, 280 whey, 53, 60, 80 WHO, 48, 283, 297, 306, 308, 311, 312, 313, 316 wide band gap, 11

wind, 114 winning, 69, 176 winter, 120, 127, 128, 129, 130, 242 wood, x, xv, 2, 12, 17, 22, 23, 25, 28, 30, 31, 32, 33, 36, 37, 41, 42, 43, 44, 63, 279, 282, 284, 285, 286, 287, 288, 289, 296 wood waste, 31 wool, 278, 339 workers, 239, 241 working conditions, 208 working hours, 314 workstation, 123 World Bank, 316 World Health Organization (WHO), 311, 334, 347 worms, 144

X xenobiotic, 59, 241, 271, 277 X-ray diffraction, 30, 83, 88, 92 X-ray diffraction (XRD), 30, 88 XRD, 30, 33, 83, 84, 87, 88, 91, 92, 96, 97, 98, 99, 102, 103 xylene, 54, 76, 77

Y yeast, 73, 158, 180, 181, 185, 191, 193, 213, 321 Yemen, 280, 300 yield, x, xiii, 2, 36, 123, 142, 158, 182, 184, 185, 186, 187, 188, 189, 190, 191, 192, 235, 238, 269, 284, 296, 302

Z Zambezi, 300 zeta potential, 2, 3, 4, 5, 22, 31 zinc, 32, 65, 66, 67, 79, 114, 122, 237 zirconium, 65 Zn, 52, 85, 86, 127, 239, 290