Environmental Biodegradation Research Focus

Environmental Biodegradation Research Focus

Wang

ENVIRONMENTAL BIODEGRADATION RESEARCH FOCUS

ENVIRONMENTAL BIODEGRADATION RESEARCH FOCUS

B.Y. WANG Editor

Nova Science Publishers, Inc. New York

Copyright © 2007 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 Environmental biodegradation research focus / B.Y. Wang (editor). p. cm. Includes index. ISBN-13: 978-60692-562-1 1. Biodegradation. I. Wang, B. Y. QH530.5.E58 2008 628.5--dc22 2007031240

Published by Nova Science Publishers, Inc.

New York

CONTENTS Preface Chapter 1

Chapter 2

Chapter 3

Chapter 4

Chapter 5

Chapter 6

Chapter 7

Chapter 8

vii Biodegradation of Polysaccharide Sourced from Virulence Factor or Plant and Pathogenic Cell Wall Constituent and its Application in Management of Phytopathogenic Disease Xianzhen Li and Xiaoyi Chen Biodegradation or Metabolism of Bisphenol A in the Environment Jeong-Hun Kang and Yoshiki Katayama From Planting to Harvest: Environmental Dissipation of the Herbicide Molinate and Proposal of a Clean-Up Methodology Célia M. Manaia and Olga C. Nunes

1

49

77

Natural Attenuation of High Concentrations of Organic Pollutants by Biodegradation in Soils L. Reijnders

101

Microbial Polyaromatic Hydrocarbon (PAH) Biodegradation in Submerged Sediment Environments Yinjie J. Tang and James Carothers

127

Microbial Degradation of 2-Benzothialzole Derivatives: A Review A. Bunescu, P. Besse-Hoggan, M. Sancelme, A. Cincilei, G. Mailhot and A.-M. Delort

159

Biodegradable Aliphatic Polyesters Derived from 1,3-Propanediol: Current Status and Promises George Z. Papageorgiou and Dimitrios N. Bikiaris

189

Aerobic Biodegradation of Fish-Meal Wastewater from Lab Scale to Large Scale Joong Kyun Kim and Geon Lee

217

vi Chapter 9

Chapter 10

Chapter 11

Chapter 12

Index

Contents Methods in Study of Biodegradation of Water Insoluble Polymer Materials Marek Koutny and Anne-Marie Delort Biodegradable Synthetic Octacalcium Phosphate Bone Substitute Takahisa Anada, Hideki Imaizumi, Shinji Kamakura and Osamu Suzuki

239

259

Biodegradation of Phenol and Resorcinol by a Halotolerant Penicillium Ana Lúcia Leitão

273

Kinetics and Metabolic Pathway of Melatonin Biodegradation by a Bacterium Isolated from the Mangrove Sediment Xiang-Rong Xu, Hua-Bin Li, Ji-Dong Gu and Xiao-Yan Li

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303

PREFACE This new book is devoted to leading-edge research on environmental biodegradation which is the destruction of organic compounds by microorganisms. Microorganisms, particularly bacteria, are responsible for the decomposition of both natural and synthetic organic compounds in nature. Mineralization results in complete conversion of a compound to its inorganic mineral constituents (for example, carbon dioxide from carbon, sulfate or sulfide from organic sulfur, nitrate or ammonium from organic nitrogen, phosphate from organophosphates, or chloride from organochlorine). Since carbon comprises the greatest mass of organic compounds, mineralization can be considered in terms of CO2 evolution. Radioactive carbon-14 (14C) isotopes enable scientists to distinguish between mineralization arising from contaminants and soil organic matter. However, mineralization of any compound is never 100% because some of it (10–40% of the total amount degraded) is incorporated into the cell mass or products that become part of the amorphous soil organic matter, commonly referred to as humus. Thus, biodegradation comprises mineralization and conversion to innocuous products, namely biomass and humus. Primary biodegradation is more limited in scope and refers to the disappearance of the compound as a result of its biotransformation to another product.Compounds that are readily biodegradable are generally utilized as growth substrates by single microorganisms. Many of the components of petroleum products (and frequent ground-water contaminants), such as benzene, toluene, ethylbenzene, and xylene, are utilized by many genera of bacteria as sole carbon sources for growth and energy. Chapter 1 - Plant cells can initiate own defense reactions to resist plant diseases on attacked by phytopathogen, in which the infection will not proceed further if such responds occur in a timely manner. Therefore the thoughtful application of the plant defense mechanisms will help plant more effectively protect against pathogen infection. Some oligosaccharides have been demonstrated to be elicitor- or antimicrobe-active. Most of these active oligosaccharides are degraded enzymatically from polysaccharide sourced from the structural constituents of plant or fungal cell walls, as well as exopolysaccharide of virulence factor of pathogens. The elicitor and antimicrobial activity greatly depends on molecular weight or degree of polymer, charge distribution, branch form, terminal groups, etc. of oligosaccharide molecules. The well-known oligosaccharides include β-glucan oligosaccharides, chitooligosaccharides, oligogalacturonides, xyloglucan-derived oligosaccharides, oligoguluronates, xanthooligosaccharide, and alginooligosaccharide. The production of oligosaccharides can be performed by both chemical and enzymatic methods, whereas the enzymatic degradation of polysaccharides is beneficial to the preparation of the

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active oligosaccharides. The enzymes involved in the degradation of polysaccharides are dependent on the carbohydrates being depolymerised and requirements to the structural feature of end products. The enzymatic degradation of polysaccharides can be managed artificially to form specific products, but the chemical degradation cannot be controlled and the hydrolysates of polysaccharides usually are their constituent units. In this chapter the biodegradation of polysaccharides including glucan, chitin/chitosan, pectin, carrageenan, xylan, xyloglucan, xanthan and alginate has been discussed. The process for polysaccharide degradation was evaluated, such as enzymes and its sources, specificity of the enzymes, enzymatic route for degradation, depolymerization and its effect on the oligosaccharide nature, preparation of bioactive oligosaccharide. The roles of oligosaccharides in plant disease resistance were also discussed generally in this chapter. Chapter 2 - Recently, there has been increasing interest in the effects of endocrine disruptors on organisms. Bisphenol A (BPA; 2,2-bis(4-hydroxyphenyl)propane; CAS Registry No. 80-05-7) is an endocrine disruptor with estrogenic activity and acute toxicity to aquatic organisms. BPA is made by combining acetone and phenol and is used mainly as a material for the production of epoxy resins and polycarbonate plastics. Due to intensified usage of these products, exposure of organisms to BPA via several routes, such as the environment and the food chain, has increased. BPA contamination in the environment occurs through several routes, such as migration from human wastes and effluent from wastewater treatment plants. BPA exposed to the environment can be biodegraded or metabolized by microorganisms (bacteria, fungi and plankton), plants, invertebrates and vertebrates (fish, amphibians and mammals). Biodegradation or metabolism is a very important step for removing or detoxifying BPA in the environment or organisms. Although some metabolites of BPA may exhibit enhanced estrogenicity or toxicity, in general, BPA biodegradation or metabolism by organisms leads to detoxication of BPA. However, excessive BPA doses cause bioaccumulation if detoxification pathways are saturated. In this chapter the authors describe 1) contamination routes of BPA, 2) biodegradation or metabolism of BPA by organisms, and 3) bioaccumulation of BPA in organisms, with the main subject of this chapter being the biodegradation or metabolism of BPA by organisms. Chapter 3 - The deliberated application and intensive use of pesticides in the environment has been leading to the contamination of air, soils, surface and ground water, and living organisms. The environmental contamination of the trophic chain with pesticides has serious negative impacts on the biological diversity and possible implications on the public health. The monitoring of environmental contamination with pesticides and the implementation of decontamination processes may contribute to minimize the impact of intensive agricultural practices. This study was conducted in a rice field situated in central Portugal, where molinate is supposed to contaminate surface and underground waters. Molinate content was monitored in water samples collected before, during and after molinate application and the results showed that molinate was dissipated in the environment, reaching concentrations of 3.9 µg l-1 in underground water and 15.8 µg l-1 in the river receiving tail waters. The feasibility of clean-up methodologies based on adsorption and/or biodegradation processes to remove molinate from these waters was assessed. At a laboratory scale, these clean-up processes led to reductions of the molinate concentration to values close to the legally recommended limits (< 2 µg l-1). Given the inability of the autochthonous microbiota to degrade molinate contaminating the agriculture effluents, the implementation of a biodegradation process requires the use of

Preface

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an exogenous molinate degrading culture. The aerobic biodegradation of molinate in the rice field waters was assayed using a defined mixed bacterial culture, previously isolated from an industrial effluent. This mixed culture is able to mineralize molinate under a wide range of operating conditions removing between 55 % (1 mM molinate at 15 °C) and 80 % (1 mM molinate and complex nutrients at 30 °C) of the initial total dissolved organic carbon. Given the low concentrations, and hence the low bioavailability, of molinate in agriculture effluents, the use of an adsorption step was considered a valuable auxiliary tool to improve the clean-up of contaminated waters. Resin Amberlite XAD-4 and activated carbon showed efficient molinate removal. The bio-regeneration of these materials, using the above mentioned mixed culture, permits the decontamination of the adsorbents, both for future reuse or for final disposal. Chapter 4 - In view of its relatively low cost, monitored natural attenuation by biodegradation is increasingly relied upon to clean up pollution of soils caused by landfills, industrial activities and major transport and storage related spills. Current policy tends to aim at reducing soil pollution to levels reflecting tolerable risk for specific recipients within a reasonable time frame. Natural attenuation by biodegradation of high concentrations of major pollutants that preferentially partition to the particle fraction of soils tends to be poor. These pollutants include hazardous hydrocarbons, highly halogenated hydrocarbons and nitroorganics. Relying on natural attenuation of these compounds, when feasible at all, leads to exceeding tolerable risk levels for a long time. Better perspectives exist for natural attenuation by biodegradation of pollutants that partition to a significant extent to the aqueous phase, especially for low molecular weight organic aromatics and chlorinated solvents. Dependent on conditions in the aquifer and the presence of suitable micro-organisms, there can be substantial biodegradation of a variety of hydrocarbons, organochlorines and oxygenates. In some cases natural attenuation has been found to bring down high levels of pollution to levels meeting current standards of tolerable risk. Predictions whether in the future levels reflecting tolerable risk can be attained, are uncertain. Uncertainty is especially large in case of expanding plumes. There are uncertainties that beset current modelling to predict future concentrations and there is uncertainty about what in the future will be considered tolerable risk. Even when it is supposed that current standards will be applied indefinitely, uncertainties related to present modelling and sampling often prevent certainty that these standards will be met in the future. In practice, natural attenuation is often falling short of attaining promised outcomes. This means that often interventions aimed at enhanced remediation will be necessary to achieve tolerable risk within a reasonable time frame. These may include enhanced biodegradation. Chapter 5 - Polycyclic aromatic hydrocarbons (PAHs) in submerged sediments can have potentially carcinogenic effects on human health through the food chain. PAH compounds persist in submerged sediment because of their very low aqueous solubilities, tendencies to adhere to sediment particles and recalcitrance to biodegradation. This chapter covers research topics important for understanding PAH bioremediation in submerged sediments: 1. microbial processes under anaerobic or aerobic sediment conditions; 2. the effects of adding inexpensive environmentally benign substances to stimulate biodegradation or improve PAH bioavailability; 3. methods to characterize microbial, chemical and physical properties in sediment sites; 4. A mathematical model linking the understanding of chemical, physical, and biological activities occurring in the sediment field. Physical capping is frequently used to treat PAH contaminated submerged sediment sites when the site is large volume and has a

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low contaminant concentration. Recent research suggests that including a physical cap with in situ bioremediation can reduce the volume of cap required to secure a site and provides a potentially economical way to rapidly remediate contaminated sediment sites. Chapter 6 - This review is focused on one particular family of pollutants, 2-benzothiazole derivatives. This group of xenobiotics containing a benzene ring fused with a thiazole ring is manufactured worldwide. After a short presentation of benzothiazole structures and their industrial applications, the fate of benzothiazoles in the environment is described both in natural waters and in wastewater treatment plants. Then data available on the toxicity of benzothiazoles are reported. The main part of this review is devoted to the microbial degradation of these compounds: i) using activated sludge and mixed cultures, ii) in soils, iii) using pure cultures. In that later case, detailed pathways of biodegradation are described for benzothiazole, 2-hydroxybenzothiazole, 2-mercaptobenzothiazole, 2-aminobenzothiazole and methabenzthiazuron. Special attention is made on methodology used to establish these pathways, namely Nuclear Magnetic Resonance (NMR). Finally photodegradation processes are described because molecular mechanisms are often closely related to those of biodegradation processes and lead to common products in the environment. To conclude, the possibility of combining these two approaches is discussed. Chapter 7 - Among biodegradable polymers, polyesters derived from aliphatic dicarboxylic acids and diols are of special importance. Polyesters of 1,3-propanediol were overlooked till recently, since the specific monomer was not available in quantities and price that might enable production of polymers. However, in recent years more attractive processes have been developed for the production of 1,3-propanediol from renewable resources. Nowadays, research on biodegradable poly(1,3-propylene alkanedioate)s, such as poly(propylene succinate) (PPSu), poly(propylene adipate) (PPAd) and poly(propylene sebacate) (PPSe), has gained an increasing interest, due to their fast biodegradation rates and their potential uses in biomedical or pharmaceutical applications, such as drug delivery systems. The odd number of methylene units in the diol segment is responsible for the lower melting points, lower degree of crystallinity and higher biodegradation rates of the specific polymers compared with their homologues based on ethylene-glycol or 1,4-butanediol. In this chapter synthesis and properties of the 1,3-propanediol based aliphatic polyesters and especially their biodegradation characteristics are reviewed. Specific attention has been paid to preparation of related copolymers and blends with other important polymers, since these techniques may offer routes for optimizing properties and produce tailor-made materials. Copolymerization of 1,3-propanediol with mixtures of aliphatic or even aromatic acids, leads to linear polyesters with improved or balanced biodegradation and mechanical properties. Blends with other biodegradable polymers have been studied recently. Finally, potential pharmaceutical applications of poly(1,3-propylene alkanedioate)s as solubilizing and stabilizing carriers for drugs are exemplified. Chapter 9 - Increasing waste disposal problems from polymer wrapping materials have resulted in constant endeavors to replace recalcitrant materials with biodegradable alternatives. The biodegradability of these materials can often simply be limited, or the processes involved are relatively slow, causing complications when applying standard methodologies and thus promoting the development of customized testing protocols. Moreover, some properties of these materials, especially their water insolubility, require further adaptations to conventional methods. This chapter brings together data found in literature, along with the personal findings and experiences of the authors. A broad variety of

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experimental methods has been used and described in literature, which are dependent on authors’ expertise or the availability of particular techniques in their laboratory. Here, a comprehensive overview of current and the most prominent techniques is provided. These include spectroscopic techniques to examine changes in the material; NMR, MS and separation methods for investigating compounds released from the material; electron microscopy methods; optical fluorescence microscopy that enables surfaces of the material and eventual microbial colonization to be visualized; different methods for biomass quantification and indicators of metabolic activity; and various ways of monitoring carbon dioxide production. The relevance of the methods for studying biodegradable synthetic polymer materials was analyzed, compared and then critically evaluated. Original microphotography and original data have been introduced to illustrate the text. Chapter 10 - Octacalcium phosphate (OCP) has been advocated to be a precursor of biological apatite crystals in bones and teeth. In fact, several studies using physical techniques demonstrated that OCP is involved as a transitory intermediate phase to biological apatite crystals in enamel, dentine and bone. The authors previous studies demonstrated that synthetic OCP facilitates bone regeneration, compared to synthetic hydroxyapatite (HA), including non-sintered stoichiometric or non-stoichiometric HA, and sintered HA ceramic, when implanted in murine and rabbit bone defects. Synthetic OCP can be replaced with newly formed bone in conjunction with its simultaneous biodegradation. Crystallographic OCP-apatite conversion of the implanted OCP advances gradually during the bone regeneration. OCP is known to be a thermodynamically less stable salt under physiological condition than HA and β-tricalcium phosphate (β-TCP); the latter is a well known biodegradable bone substitute ceramic. The biodegradability predicted from the solubility isotherm is actually reproduced in the implantation of these synthetic calcium phosphate compounds into bone defects. OCP-apatite conversion induces various physicochemical reactions on the crystal surfaces, including bio-molecule adsorption, and may be involved in the stimulatory effect to enhance osteoblastic cell differentiation in vitro and bone formation in vivo. Chapter 11 - Many industries are known to generate wastewater enriched in phenolic compounds. These include petrochemicals, basic organic chemical manufacture, coal refining, pharmaceutical and tanning. Consequently, these compounds are commonly encountered in industrial effluents and surface water. Due to its high toxicity as shown by ecotoxicological studies, several methods have been reported for the removal of these pollutants from wastewater. Additionally to this toxicity problem some of these industrial effluents are likely to generate highly saline wastewaters. The discharge of such wastewaters containing at the same time phenol and phenolic compounds and high salinity without prior treatment is known to negatively affect the aquatic life, agriculture and potable water. Biological treatment with halotolerant/halophilic microorganisms is considered advantageous over the other physical and chemical methods as it leads to complete mineralization of phenolic compounds, is one of the safest, least costly and most socially acceptable. Halotolerant microorganisms are well known for their great versatility to remove pollutants, under saline and non saline conditions. Penicillium chrysogenum is an economically important ascomycete used as producer of penicillin. However, little attention has been paid to the ability of this microorganism to transform or metabolize compounds that are pollutants. This article presents a different approach to a classic problem. It employs an individual test on marine organism of trophic level 2 to validate bioremediation process by halotolerant fungus.

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A biodegradation process is an effective method of remediation if the toxicity of the system decreases. The purpose of this study was to compare cell growth and biodegradation of single phenol and resorcinol at high initial substrate levels by halotolerant strain, P. chrysogenum CLONA2, under saline and non saline conditions, adapted with either phenol or resorcinol, in a batch system, and to study the inhibitory and enhanced effect during the biodegradation of phenol and resorcinol. Single and binary substrate experiments were performed. HPLC analysis shows that halotolerant strain, Penicillium chrysogenum CLONA2, degraded up to 300 mg/l of both xenobiotics compounds in mineral salts medium with 58.5 g/l of sodium chloride. When phenol and resorcinol were together in low concentrations (< 15 mg/l and < 30 mg/l, respectively), phenol enhanced resorcinol degradation. P. chrysogenum CLONA2 metabolized phenol faster than resorcinol when present as the sole carbon source. The acute toxicity of phenol and resorcinol, individually and in combination, to larvae of the Artemia franciscana has been verified after and before bioremediation process with P. chrysogenum CLONA2. Resorcinol was more toxic than phenol. The authors findings indicate that mixtures of resorcinol and phenol had an effect more toxic in A. franciscana than individual phenolic compounds. Chapter 12 - Melatonin (MLT) is a hormone produced primarily by the pineal gland. It can be found in animals and humans as well as a number of bacteria, fungi and plants. Since it has been widely used as the healthcare product in the world market, MLT entered the various environmental compartments by different ways. Recent research found that MLT may suppress the production of testosterone, decrease semen quality, and affect sexual activity and reproduction of animal and human. The fate of melatonin in environment has attracted increasing public and scientific concerns in recent years. In this chapter, biodegradability of MLT was studied for the first time. A strain that can efficiently degraded melatonin was isolated from the mangrove sediment. This strain, a gram-negative bacterium identified as Shewanella putrefaciens, can grow on melatonin as sole sources of carbon and energy under aerobic conditions. The growth was greatly enhanced by the addition of a small amount of yeast extract. Effects of melatonin concentration, pH, temperature and salinity on MLT biodegradation were studied, respectively. The experimental results showed that 50 mg l-1 melatonin could be degraded within 2 d under the optimal condition (pH 7.0, salinity 15‰ and temperature at 37 °C). The process of MLT biodegradation was monitored by highperformance liquid chromatography with ultra-violet detection. The biodegradation of melatonin could be fitted to a first-order kinetic model. The major metabolites of melatonin biodegradation were identified by high-performance liquid chromatography and gas chromatography-mass spectrometry, and a preliminary metabolic pathway of melatonin was proposed. The results obtained are helpful to understand environmental behavior of MLT, and also could be used for the bioremediation of MLT-contaminated site, such as the wetland of the Mai Po Natural Reserve in Hong Kong.

In: Environmental Biodegradation Research Focus Editor: B. Y. Wang, pp. 1-47

ISBN: 978-1-60021-904-7 © 2007 Nova Science Publishers, Inc.

Chapter 1

BIODEGRADATION OF POLYSACCHARIDE SOURCED FROM VIRULENCE FACTOR OR PLANT AND PATHOGENIC CELL WALL CONSTITUENT AND ITS APPLICATION IN MANAGEMENT OF PHYTOPATHOGENIC DISEASE Xianzhen Li∗ and Xiaoyi Chen Department of Bio & Food Engineering, Dalian College of Light Industry

1. ABSTRACT Plant cells can initiate own defense reactions to resist plant diseases on attacked by phytopathogen, in which the infection will not proceed further if such responds occur in a timely manner. Therefore the thoughtful application of the plant defense mechanisms will help plant more effectively protect against pathogen infection. Some oligosaccharides have been demonstrated to be elicitor- or antimicrobe-active. Most of these active oligosaccharides are degraded enzymatically from polysaccharide sourced from the structural constituents of plant or fungal cell walls, as well as exopolysaccharide of virulence factor of pathogens. The elicitor and antimicrobial activity greatly depends on molecular weight or degree of polymer, charge distribution, branch form, terminal groups, etc. of oligosaccharide molecules. The well-known oligosaccharides include βglucan oligosaccharides, chitooligosaccharides, oligogalacturonides, xyloglucan-derived oligosaccharides, oligoguluronates, xanthooligosaccharide, and alginooligosaccharide. The production of oligosaccharides can be performed by both chemical and enzymatic methods, whereas the enzymatic degradation of polysaccharides is beneficial to the preparation of the active oligosaccharides. The enzymes involved in the degradation of polysaccharides are dependent on the carbohydrates being depolymerised and

2

Xianzhen Li and Xiaoyi Chen requirements to the structural feature of end products. The enzymatic degradation of polysaccharides can be managed artificially to form specific products, but the chemical degradation cannot be controlled and the hydrolysates of polysaccharides usually are their constituent units. In this chapter the biodegradation of polysaccharides including glucan, chitin/chitosan, pectin, carrageenan, xylan, xyloglucan, xanthan and alginate has been discussed. The process for polysaccharide degradation was evaluated, such as enzymes and its sources, specificity of the enzymes, enzymatic route for degradation, depolymerization and its effect on the oligosaccharide nature, preparation of bioactive oligosaccharide. The roles of oligosaccharides in plant disease resistance were also discussed generally in this chapter.

2. INTRODUCTION The chemical pesticides have successfully been used to control plant diseases due to their quick and effective management. However their incessant and indiscriminate use will have harmful effects on human health and environment. In recent years a large number of synthetic pesticides have been banned because of their undesirable attributes such as high and acute toxicity, long degradation periods, accumulation in food chain and an extension of their power to destroy both useful and harmful pests [Annon, 1996; Barnard et al., 1997]. Besides, many pathogenic microorganisms and insect pests have developed resistance against chemical pesticides, which seriously hinders the management of diseases of plant [May, 1985; Williams and Heymann, 1998]. Therefore, an urgent requirement for searching biological control agents to manage phytopathogen has been intensified in recent years. Plants are capable of initiate various defense reactions, such as the production of phytoalexins, the expression of antimicrobial proteins, the generation of the reactive oxygen species, and the reinforcement of cell walls, when they are attacked by pathogens. Therefore it is believed that the thoughtful application of the plant own defense mechanisms can lead to more effective protection against phytopathogens [Somssich and Hahlbrock, 1998]. Carbohydrates are the most abundant materials among the polymers in nature, which have been thought only to play roles as energy storage molecules or structural elements in cell walls for a long time. However, they have recently been actively studied as important biological macromolecules due to its great structural diversity for encoding biological information. Oligosaccharide signals have been shown to regulate the defensive or symbiotic processes in plants [Coté and Hahn, 1994; Darvill et al., 1992], and possess versatile functional properties such as antitumor activity, immune-enhancing effects, antimicrobial activity, and growth stimulation on probiotics [Hirano and Nagao, 1989; Lee et al., 2002; Suzuki, 1996; Suzuki et al., 1986]. An oligosaccharide is any short chain of sugar residues interconnected by glycosidic linkages. Oligosaccharides can be produced artificially by the enzymatic fragmentation of polysaccharides. Some oligosaccharides have been demonstrated to be elicitor- or antimicrobe-active, which can stimulate plant defense responses and may help plants resist infective disease [Shibuya and Minami, 2001]. ∗

Correspondence should be sent to Dr. Xianzhen Li, Department of Bio & Food Engineering, Dalian College of Light Industry, Ganjing Qu, Dalian 116034, PR CHINA; Tel: (86) 411 86314195; Fax: (86) 411-86323646; Email: [email protected]

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3

The first bioactive oligosaccharide was degraded from the fungal cell walls of phytopathogen Phytophthora sojae, which was able to induce the synthesis of phytoalexins in plant cells to prevent fungal infection [Sharp et al., 1984a]. Other active oligosaccharides derived from the cell walls of plant and pathogen were subsequently produced and isolated, which have shown important roles in plant protection. Most of the known oligosaccharides identified as having elicitor or antimicrobial activity are from the structural components of fungal or plant cell walls, as well as the virulence factor of exopolysaccharides by partial degradation. The well-known active oligosaccharides include β-glucan oligosaccharides, chitin/chitosan oligosaccharides, pectic oligosaccharides, xyloglucan-derived oligosaccharides, oligoguluronates, xanthooligosaccharide, alginooligosaccharide, etc. [Fry et al., 1993b; Liu et al., 2005]. Their bioactivities are greatly dependent on the degradation degree, the charge number and distribution, and the branch form and end groups, in which only a few oligosaccharides exhibit elicitor or antimicrobial activity. In this chapter we will review the biodegradation of polysaccharides for preparation of bioactive oligosaccharides by different microorganisms and their enzymes, including enzymes and their sources for biodegradation, process for enzymatic degradation, production of bioactive oligosaccharides, oligosaccharide nature and its effect on bioactivity, function involved in plant defensive responses, roles in plant disease resistance.

3. POLYSACCHARIDE FRAGMENTS FROM FUNGAL AND PLANT CELL WALLS OR VIRULENCE FACTORS ACTED AS ELICITORS OR ANTIMICROBES Oligosaccharides released from fungal and plant cell walls or some virulence factors of phytopathogen are powerful signaling elicitors, capable of acting at very low concentrations to convey information to the plant under attack. In response to this information, plant defense response was activated, leading to the induction of genes that encode enzymes responsible for the synthesis of phytoalexins and other defensive compound. Some of the oligosaccharides with bioactivity were illustrated in Figure 1. The first oligosaccharide identified as having elicitor activity was isolated from a fungal pathogen of soybean, Phytophthora sojae [Sharp et al., 1984a]. This branched glucan heptasaccharide, composed of β-1,3-and β-1,6-linked glucose residues, is the smallest possible unit from the fungal cell wall capable of inducing the synthesis of phytoalexins by plant cells, an antimicrobial molecules for resistance on microbial infection. Subsequently, the other structural components of the cell walls of pathogenic fungi have also been shown to elicit defense responses in plants such as chitin/chitosan oligosaccharides [Coté and Hahn, 1994]. In addition, polysaccharide fragments released from plant cell walls play an important role in signaling a plant defensive response. Oligogalacturonides were released from the pectic component of plant cell walls via the enzymatic degradation with polygalacturonase and pectate lyase, which can elicit a broad spectrum of plant defense responding reactions, including the increased synthesis of phenylalanine ammonia-lyase (PAL), chalcone synthase, chitinases, β-glucanase, and protease inhibitors [Bishop et al., 1984; Coté and Hahn, 1994;

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Glc

Glc

α -1,3

β -1,6

Glc

GlcN

β -1,6 β -1,6 β -1,6 Glc Glc Glc Glc β -1,3 β -1,3 Glc Glc Glucan heptasaccharide

α -1,3

β -1,4

α -1,3 α -1,3 α -1,3 Glc Glc Glc Glc α -1,3 Glc Hexosaccharide

GlcN

β -1,4

GlcN

β -1,4

GlcN

Chitosan oligosaccharides GlcNAc

β -1,4

GlcNAc

β -1,4

GlcNAc

β -1,4

GlcNAc

Chitin oligosaccharides GlcA

α -1,4

GlcA

α -1,4

GlcA

α -1,4

GlcA

α -1,4

GlcA

α -1,4

GlcA

Pectin oligosaccharides (oligogalacturonides) β -1,4 β -1,4 β -1,4 GlcN GlcN GlcN GlcN α -1,6 α -1,6 α -1,6 Xyl Xyl Xyl α -1,2 Gal α -1,2 Fuc Xyloglucan nonasaccharide Figure 1. Structure of some bioactive oligosaccharides derived from fungal and plant cell walls. Abbreviations: Glc, glucose; GlcN, glucosamine; GlcNAc, N-acetylglucosamine; GalA, galacturonic acid; Xyl, Xylose; Gal, galactose; Fuc, fucose.

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5

Ryan and Farmer 1991]. Another oligosaccharide signal derived from plant cell walls is a xyloglucan oligosaccharide. This bioactive oligosaccharide was released from plant cell walls by cleavage of xyloglucan catalyzed by endo-β-1,4-glucanase [Fry et al., 1993a]. Xyloglucan-derived oligosaccharides were also reported to influence on defense reactions and elicit the release of the related rotting enzymes that can hydrolyze the structural constituents of the host plant further [Slováková et al., 1994]. Xanthooligosaccharide, produced by the degradation of xanthan via a series of xanthan-degrading enzymes, is another interesting oligosaccharide, which has been demonstrated to have elicitor and antibacterial activity [Liu et al., 2005; Qian et al., 2006]. Xanthan is an extracellular heteropolysaccharide produced by a phytopathogenic bacterium Xanthomonas campestris pv. campestris [Rogovin et al., 1961], which was involved in the pathogenicity caused black rot lesions to form on cruciferous plants [Daniels et al. 1987]. Similar oligosaccharide also includes alginooligosaccharide derived from alginate, a virulence factor of phytopathogen Pseudomonas syringae [PeñalozaVázquez et al., 2004].

1) Mechanism of plant defense response When plants are attacked by pathogenic microorganisms, they have the ability to initiate various defense reactions themselves against pathogenic organisms infection, such as the induction of secondary metabolic enzymes like PAL for phytoalexin formation, the reinforcement of cell walls by deposition of callose, the synthesis of hydrolytic enzymes like chitinase and β-1,3-glucanase for broken cell walls, a rapid formation of the plant hormone ethylene, and the release of reactive oxygen species [Alvarez et al., 1998; Bell, 1981;Boller, 1995; Hahlbrock and Scheel, 1989;Mehdy, 1994] (Figure 2). If these reactions occur in a timely manner, the infection will not proceed further. However, if these defense reactions occur too late or are suppressed, the infection process will proceed successfully, leading to various plant diseases happen [Somssich and Hahlbrock, 1998]. The recognition of phytopathogens by plant cells depends on the perception of elicitors generated by the pathogen. Fungal or plant cell wall fragments and molecules secreted by the pathogen can induce signaling cascades that activate a cellular response to minimize injury to host plant [Blumwald et al., 1998; Dixon et al., 1994]. It is well known that the presence of oligosaccharides on the cell membrane of pathogen can act as elicitors. And the receptors for oligosaccharide elicitors were also demonstrated to occur on plant cell membranes [Cheong and Hahn 1991; Yoshikawa et al. 1983]. Upon infection by a phytopathogen, plants can percept its invasion and commonly activate a variety of defense mechanisms. For example, the challenged plants will synthesize β-glucanase and chitinases instantly, catalyzing the partial degradation of the cell wall of plant or pathogen to form β-glucan, chitosan/chitin oligosaccharides [Flach et al., 1992]. Such oligomers released by degradation of fungal and plant cell walls acted as elicitors in turn can facilitate additional plant defense mechanisms [Frindlender et al., 1993], such as cell wall fortification [Bradley et al., 1992], defense-related gene expression [Jabs et al., 1997; Levine et al., 1994], and phytoalexins synthesis [Bradley et al., 1992; Jabs et al., 1997]. Maurhofer et al. [1994] reported that the induction of systemic resistance by Pseudomonas fluorescens was correlated

6

Xianzhen Li and Xiaoyi Chen

pathogen

pathogen

pathogen

nuclear

pathogen

pathogen

pathogen

pathogen

Plant nuclear

mRNA

1

2

pathogen

4

3

Plant mRNA

5

More defense response Pathogen Oligo-glucan Cell death

Pa tho

ge n

Chitooligosaccharide

Plant

Oligogalacturonide 4 Chitinase

1 Phytoalexin

2 Active oxygen species

3

Glucanase

5

Pecticn lyase

Figure 2. Plant defense response on infected by phytopathogen or elicitor treatment.

with the accumulation of β -1,3-glucanase and chitinase at the site of penetration of fungal hyphae. These enzymes acted on the fungal cell walls resulting in the degradation and the loss of inner contents of cells, and leading to the destabilization of the organism [Benhamou et al., 1996; Ji and Kuc, 1996]. The fungal cell wall elicitors have been reported to elicit various defense reactions in greengram [Ramanathan et al., 2000]. So the pathogen invasion can be minimized by produced elicitors in the infection process. It has been proved that there are two types of defense mechanism occurred in plant cells. One is endogenous defense mechanism as described above, which can be induced in response to the attack of pathogens. Another is exogenous defense response, which is initiated by the foreign elicitor treatment. So expression of β-glucanases and chitinases can be activated not only upon the challenge of plant tissues by microbes, but also upon exposure to certain elicitors [Kirsch et al., 1993; Kombrink and Hahlbrock, 1986; Mauch et al., 1984]. Structural polysaccharide fragments from plant or pathogen cell walls can serve as elicitors in many plant species [Lee et al., 1999]. Induction of plant defense genes by prior application of biological inducers is called induced resistance, which can prevent plants from the infective disease of pathogen or virus. It is well known that the defense genes are inducible genes and appropriate stimuli or signals are needed to activate them. In many plant species, resistance may be induced against pathogens by means of pretreatment with some of oligosaccharides. Once resistance was induced, the plant expressed a number of inducible defense responses or mechanisms that usually coincides with the accumulation of pathogenesis-related proteins [Gatz, 1997; Greenberg, 1997; Hunt and Ryals, 1996; Keen, 2000; Lawton and Lamb, 1987]. PAL is a key enzyme in the production of phenolics and phytoalexins [Daayf et al., 1997]. PAL activity could be induced in plant-pathogen interactions and exogenous elicitor treatment [Ramanathan et al., 2000]. Other defense enzymes include pathogenesis-related

Biodegradation of Polysaccharide Sourced from Virulence Factor…

7

proteins such as peroxidase, which is a key enzyme in the biosynthesis of lignin [Bruce and West, 1989]. Increased activity of cell wall bound peroxidases has been elicited in different plants such as cucumber, rice, and tomato due to pathogen infection [Chen et al., 2000; Mohan et al., 1993; Reimers et al., 1992]. Thaumatin-like proteins belong to PR-5 family, showing antifungal activity and enhancing resistance to pathogen infection [Datta et al., 1999]. Induction of defense proteins makes plant resistant to pathogen invasion [Van Loon, 1997]. Therefore the induction of plant resistance mechanisms by application of elicitors has been suggested as an alternative approach for crop disease control [Benhamou, 1996; Cartwright et al., 1977; Gatz, 1997; Hunt and Ryals, 1996].

2) Mechanism of antimicrobial activity Some oligosaccharides can bind to the microbe’s carbohydrate-binding proteins to prevent bacterial attachment on host cells or clear bacteria already attached [Zopf and Roth, 1996]. The mostly accepted mechanism to explain the possible antibacterial actions of chitooligosaccharides has been proposed by EI-Ghaouth et al. [1992]. It was asserted that chitooligosaccharide reacted with the pathogen cell surface to alter the permeability characteristics of microbial cell membrane, and further prevented the entry of materials or caused the leakage of materials, and finally led to the death of bacteria. More adsorbed chitooligosaccharide will result in greater changes in the structure of the cell wall and in the permeability of the cell membrane. Sudharshan et al. [1992] also demonstrated that the chitooligosaccharide could alter permeability and further prevent the entry of materials or cause leakage of cell constituents, leading to the death of bacteria. Another suggested mechanism for antibacterial activity of chitooligosaccharide was the blockade of RNA transcription by adsorption of penetrated chitosan to bacterial DNA [Kim et al., 2003]. The antibacterial activity of oligosaccharides is related to the hydrophilicity of cell wall, whereas the inhibition capacity does not fit well with the hydrophilicity of cell wall for Gramnegative bacteria [Chung et al., 2004]. So the hydrophilicity of the cell wall could not fully explain the difference in the antibacterial activity for Gram-negative bacteria. It was found that the distribution of negative charges on their cell surfaces was quite different although cell wall hydrophilicity was similar among Gram-negative bacteria. Hence, positively charged chitooligosaccharide had higher antibacterial activity or inhibition activity in Staphylococcus aureus than Streptococcus faecalis [Chen et al., 2002]. This clearly explained why most Gram-negative bacteria were sensitive to chitooligosaccharide, and negative charge density on the cell surface apparently determined whether the bacteria were easily inhibited by chitooligosaccharide or not. Because the amount of adsorbed chitooligosaccharide to the different bacterial cells is exactly the same order determined for the antibacterial activity of chitooligosaccharide on these bacteria, the antibacterial activity of chitooligosaccharide and the surface characteristics of the cell wall are closely related [Wang, 1992]. More negatively charged cell surfaces had a greater interaction with oligosaccharides. The number of amino groups in chitooligosaccharides has been proved to play a major role in antibacterial activity, and several mechanisms have been proposed to describe this activity [Chen et al., 2002]. In general, when the positively charged group of chitooligomers interacts with the negatively charged carboxylic acid group of macromolecules on bacterial cell surface, the polyelectrolyte complexes will be formed on the cell surface of pathogen

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Xianzhen Li and Xiaoyi Chen

[Choi et al., 2001; Kim et al., 2003]. This complex acts as an impermeable layer around the cells and suppresses the metabolic activity of the bacteria by blocking of nutrient permeation through the cell wall, and lastly results in the cell death. With respect to the formation of polyelectrolyte complexes, higher number of primary amino groups presented in chitooligosaccharide can make stronger interactions with bacterial cells [Tsai et al., 2002]. Many pathogens use carbohydrate-binding proteins to attach to cells and initiate plant disease. The first line of defense against these infectious diseases consists of decoy oligosaccharides in the mucous layer. When decoy oligosaccharides bind to the microbe’s carbohydrate-binding proteins competitively, the attached pathogens are released from the host cells and cleared by the physiological mechanism (Figure 3). So the soluble oligosaccharides can both prevent bacterial attachment and separate bacteria already attached [Zopf and Roth, 1996]. The decoy is a homologue, regardless of size, provided it is a numerically and isomerically identical to the corresponding fragment of the native carbohydrate. The pathogen proteins (adhesins, lectins) have strict requirements for their oligosaccharide ligands. Usually the specific sugar sequence required by an adhesin is at the terminal, non-reducing end of the oligosaccharide chain, although many adhesins also recognize internal sugars [Zopf and Roth, 1996].

a

Pathogen

Plant

b

Plant

Pathogen

Plant cell Figure 3. Binding of cell surface carbohydrate on microbial adhesion (a) and carbohydrate-binding proteins on bacterial cell surface occupied by oligosaccharides preventing their interaction with plant cell surface (b). Adapted and modified from Zopf and Roth [1996].

A lock-and-key hypothesis was proposed for explaining the interrelation between the plant and pathogen [Zopf and Roth 1996]. The cell surface carbohydrates are named as "keys", and carbohydrate binding proteins on cell membrane of pathogen are asserted as the

Biodegradation of Polysaccharide Sourced from Virulence Factor…

9

"locks" in such lock-and-key interactions. The role of oligosaccharides is to make themselves to excellently competitive inhibit the target interaction. The optimum size of the competing oligosaccharide depends on the number of monosaccharide subunits recognized by the complementary protein. If the carbohydrate-binding domain of a protein has a cleft fitting four terminal residues of a specific octasaccharide, for example, the terminal tetrasaccharide alone will bind to the protein with an avidity being the same as that of the native octasaccharide. Whereas the terminal trisaccharide will bind less well than the native compound, and the terminal disaccharide worse still. A non-carbohydrate analogue may bind efficiently to a microbial adhesin [Ofek et al., 1990], but it is more likely to be toxic and immunogenic than that of the carbohydrate homologue. When cell surface carbohydrates are linked to a flexible polymer backbone to create a macromolecule, they can simultaneously engage many adhesin molecules at a bacterial cell surface, forming a stable complex (Figure 3a). Whereas the oligosaccharides may effectively competitive inhibit the combination between plant and pathogen. The pathogen attached on plant cells can be removed on binding of oligosaccharide to a few sites of microbial adhesin (figure 3b) [Spaltenstein and Whitesides, 1991].

4. BIODEGRADATION OF POLYSACCHARIDES 1) Biodegradation of glucan Certain plant defense reactions are elicited by compounds referred to as elicitors, such as oligosaccharides released from fungal and plant cell walls [Yoshikawa et al., 1993]. Oligo-βglucan elicitor, released from the cell wall of the phytopathogenic fungus Phytophthora megasperma by soybean glucanases, caused defense reactions in soybean cells [Umemoto et al., 1997]. The first event between the soybean and fungus is an attack of the fungal cell wall by soybean β-1,3-glucanase, resulting in the release of active oligo-β-glucan elicitors. The released oligo-β-glucan then initiated phytoalexin accumulation in plants [Yoshikawa et al., 1981]. β-Glucans were recognized to be actively involved in plant-pathogen interactions in the mid-1970s, the ability of which to induce phytoalexin accumulation in soybean cells was first detected in the culture filtrates of Phytophthora sojae [Ayers et al., 1976a]. These elicitors are composed of 3-, 6-, and 3,6-linked β-glucosyl residues, a composition very similar to glucans that are major constituents of various mycelial walls [Ayers et al., 1976b; Bartnicki-Garcia, 1968]. The elicitor-active glucans can be released from the mycelial walls of Phytophthora spp. by either partial chemical degradation or enzymatic hydrolysis [Ayers et al., 1976b; Keen and Yoshikawa, 1983; Sharp et al., 1984a; Yoshikawa et al., 1981]. Lichenase cleaved the β-(1,4)-linkages adjacent to α-(1,3)-linkage. The oligosaccharides thus obtained were the (1,4)-linked building blocks of β-glucan with (1,3)-linked end group [Johansson et al., 2000; Wood et al., 1994]. Other enzymes used for depolymerization of βglucan samples were cellulase, which cleaved only β-(1,4)-glycosidic linkages [Roubroeks et al., 2001], and β-glucosidase together with lichenase were used for quantitative measurement of β-glucan in cereal products [Boyac et al., 2002; McCleary and Codd, 1991].

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Xianzhen Li and Xiaoyi Chen

Eight hexa (β-D-glucopyranosyl)-D-glucitols, two of which were not separated from each other, were purified to homogeneity from the mixture of oligoglucosides generated by partial hydrolyzed mycelial walls of Phytophthora megaaperma f. sp. glyeinea [Sharp et al., 1984b]. Only one of these hexa (β-D-glucosyl)-D-glucitols was shown to have elicitor activity. The smallest elicitor-active oligoglucoside is a branched hepta-β-glucoside with a backbone of five (1,6)-linked β-glucosyl residues and have two terminal β-glucosyl residues attached at C3 of the second and fourth backbone glucose rings [Sharp et al., 1984a]. The structure of this hexa (β-D-glucopyranosyl)-D-glucitol was shown as below: β D Glcp (1

6) β D Glcp (1 3 1 β D Glcp

6) β D Glcp (1

6) β D Glcp (1 3

6) Glucitol

1 β D Glcp

The glycosyl-linkage compositions of the hexa (β-D-glucosy1)-D-glucitols indicated that only small structural differences existed between the elicitor-active and the elicitor-inactive molecules. These results indicated that the similarity in structure of six elicitor-inactive oligosaccharides to a highly defined structure of the elicitor-active molecules established was required for elicitor activity. The elicitor-active hexa (β-D-glucopyranosy1)-D-glucitol was the first example of complex carbohydrate acting as a regulatory molecule in plants [Sharp et al., 1984a]. Evidences of glucan as an elicitor were obtained from the crude or only partially purified fungal cell wall fractions, such as the Phytophthora sojae cell wall hydrolysate, a heterogeneous β-1,3-1,6 glucan extracted from the mycelial walls of Phytophthora sojae f. sp. glycinea [Sharp et al., 1984a]. The elicitor activity of this glucan was mainly studied in leguminous plants but it was also reported to induce a Gly-rich protein [Brady et al., 1993] and antiviral protection in tobacco cells [Kopp et al., 1989]. A pure glucan heptasaccharide [Sharp et al., 1984a] prepared from Phytophthora sojae glucan induced the synthesis of phytoalexins in soybean (Glycine max) [Sharp et al., 1984a]. The minimal structural requirements for the elicitation of phytoalexin synthesis in soybean by this glucan were established as a succession of five β-1,6-linked glucosyl residues with two side branches of β1,3-glucose [Cheong et al., 1991]. Specific binding sites for the β-1,6-1,3 heptaglucan from Phytophthora sojae have been described in soybean [Cheong et al., 1991], alfalfa (Medicago sativa), bean (Phaesoleus vulgarus), lupine (Lupinus albus), and pea (Pisum sativum) [Cosio et al., 1996; Côté et al., 2000]. The smallest elicitor-active β-glucan has been determined (hepta-β-glucoside), and the chemically synthesized elicitor was reported not only to induce phytoalexin accumulation but also to have a high affinity for the plasma membrane fraction [Cheong et al., 1991; Sharp et al., 1984a]. These reports strongly suggested that these oligo-glucan elicitors bind to a receptor on the plasma membrane of soybean, resulting in the accumulation of phytoalexins.

Biodegradation of Polysaccharide Sourced from Virulence Factor…

11

2) Biodegradation of chitin and chitosan Chitin is a natural polymer, being a linear polysaccharide of β-(1,4)-linked N-acetyl glucosamine (GlcNAc) residues, found in the exoskeletons of crustacea and insects and in the cell walls of certain fungi. It is the second most abundant polymer in nature next to cellulose [Rudrapatnam et al., 2003]. However the insolubility of chitin in most solvents seriously restricts its biological activities. Chitosan is derived from chitin by deacetylation in the presence of alkali, which is a copolymer consisting of β-1,4-linked glucosamine (GlcN) with various degrees of N-acetylated glucosamine residues [Arvanitoyannis et al., 1998]. Chitosan has been demonstrated to have different biological activities, such as antitumor effect , antibacterial effect, and antifungal effect [Choi et al., 2001; Jeon et al., 2001; Roller et al., 1999; Seo et al., 2000]. Like the difficult situation chitin is facing, however, the application of polymeric chitosan was also limited due to its solubility only in weakly acidified water and its high viscosity. Recently the studies on chitin and chitosan have been paid more interest in biodegradation of chitin and chitosan to form chitooligosaccharides, especially the hydrolysis of chitosan into oligosaccharides. Because the chitosan oligomers not only are water-soluble and have low viscosity due to their shorter chain lengths and free amino groups in D-glucosamine units, but also possess versatile functional properties such as antitumor activity, immuno-enhancing effects, enhancement of protective effects against infection with some pathogens, antifungal activity, and antibacterial activity [Jeon & Kim, 2001; Jeon et al., 2000; Jeon et al., 2001; Hirano & Nagao, 1989; Kendra et al., 1989; Suzuki, 1996; Suzuki et al., 1986; Tokoro et al., 1989; Tsukada et al., 1990]. With respect to antimicrobial activity, it has been known that chitosan oligomers are superior to chitin oligomers because the partially degraded chitosan possesses a lot of polycationic amines, which can interact with the negatively charged residues of macromolecules on the cell surface of microorganisms and subsequently inhibit the cell growth of microorganisms [Young & Kauss, 1983]. The antimicrobial effect of chitooligosaccharides has also been shown to be greatly dependent on their degree of polymerization (DP) or molecular weight and requires glucosamine polymers with DP 6 or greater [Kendra and Hadwiger, 1984]. Chitosan can be depolymerized by partial hydrolysis with concentrated HCl [Horowitz et al., 1957]. However, acidic hydrolysis produced low yields of oligosaccharides and a large amount of its constituent units D-glucosamine. Also, the oligosaccharides prepared by the acidic hydrolysis might be toxic because of chemical changes during conversion. Therefore, the preferred method for producing chitooligosaccharides with specific lengths and sequences should be depolymerization of chitin/chitosan by enzymatic hydrolysis [Sikorski et al., 2005]. Up to now, there have been many different enzymes being isolated and studied for this purpose. The hydrolysis of chitin/chitosan to monomer N-acetyl glucosamine and glucosamine oligosaccharides has been studied for decades, in which the main representative chitolytic enzyme was chitinase catalyzing the hydrolysis of chitin. However, it was found that the chitin degradation was difficult due to its insolubility. The better suitable substrate for chitinase should be partially deacetylated chitin. Chitin/chitosan is generally susceptible to a number of different enzymes indicating its broad substrate specificity [Aiba, 1994a, b]. Up to date, a range of chitino/chitosanolytic enzymes have been found in the most living organisms including bacteria [Lee et al., 1996; Ohtakara et al., 1990; Varum et al., 1996; Yang et al.,

12

Xianzhen Li and Xiaoyi Chen

2005], actinomycetes [Hoell et al., 2006; Ohtakara et al., 1990], fungi [Kim et al., 1998; Muzarelli et al., 1994a], plant [Eilenberg et al., 2006; Lakhtin et al., 1995], and mammals or humans [Boot et al., 2001; Escott and Adams, 1995; Renkema et al., 1995]. Chitinolytic bacteria produced multiple chitinases, but there was comparatively little information available about the properties and roles of the individual chitinases in a chitinolytic system. Alteromonas sp. strain O-7 secretes four chitinases (ChiA, ChiB, ChiC, and ChiD) in the presence of chitin, in which ChiA plays a central role in chitin degradation [Orikoshi et al., 2005]. Amycolatopsis orientalis (Nocardia orientalis) was known to secrete various chitinolytic enzymes such as chitinase, β-N-acetyl hexosaminidase, and endo-βglucosaminidase [Nanjo et al., 1989, 1990]. These enzymes have been shown to catalyze efficient transglycosylation activities [Usui et al., 1987, 1990]. In addition, some other common carbohydrases and proteases have also been proved their hydrolytic ability on chitosan to produce chitooligosaccharides with different molecular weights [Aiba, 1994a; Zhang et al., 1999]. Lysozyme from hens’ eggs has been investigated and shown to be the most efficacious when the chitosan is only partially deacetylated [Nordtveit et al., 1994, 1996]. Enzymes from a variety of sources have been used for examining chitosan degradation. Uchida et al. [1989] reported that chitosanase from Bacillus sp. produced mainly the oligosaccharides with DP 2–6 and a small amount of D-glucosamine after prolonged incubation, suggesting its endo-action on chitosan and its ability to degrade partially chitosan molecules. The susceptibility of chitosan to a number of different enzymes has been investigated. Aiba [1994a, 1994b] carried out the hydrolysis of partially N-acetylated chitosan with chitinase and lysozyme because these enzymes can recognize N-acetyl glucosamine residues in chitosan. Pantaleone et al. [1992] reported the hydrolytic susceptibility of chitosan to a wide range of enzymes, including 10 kinds of glycanases, 21 kinds of protease, five lipases and a tannase, which were derived from different bacterial, fungal, mammalian and plant sources. Among them papain from Carica papaya and hemicellulase and lipase from Aspergillus niger were reported as effective enzymes to hydrolyze chitosan. Muzzarelli et al. [1994b, 1995] examined the action of papain and lipase in depolymerizing chitosan. From these results, a lot of commercial enzymes have been developed for efficient hydrolysis of chitosan. These enzymes, however, were added at relatively high concentrations, while chitosanase showed substantial activities at low concentrations. Further, it has been revealed that the structure of glycosidic bonds in chitosan affects enzymatic hydrolysis process. Differentially deacetylated chitosan have four different types of randomly distributed glycosidic bonds in their structures. These include linkages GlcNGlcN, GlcNAc-GlcN, GlcN-GlcNAc and GlcNAc-GlcNAc. Egg white lysozyme was found to be almost exclusive towards the cleavage of glycosidic linkage of GlcNAc-GlcNAc, while Bacillus chitosanase was found to be highly specific towards GlcN-GlcN linkages [Varum et al., 1996]. In addition, chitinase could act on partially N-acetylated chitosan by recognizing GlcNAc residues in the chitosan sequence [Aiba, 1994b]. Chitosanases from different organisms also differ in their catalytic action and that is mainly dependent on deacetylated degree of chitosan [Kurita, 1998]. However, it has been generally observed that chitosanases obtained from microbes produce relatively a higher yield of chitooligosaccharides compared to chitosanases from the other sources. Although microbial chitosanases have shown to have excellent performances in chitooligosaccharides production, they are too expensive to be utilized in large-scale

Biodegradation of Polysaccharide Sourced from Virulence Factor…

13

industrial applications. Therefore, other commercial enzymes were utilized under specific conditions to produce chitooligosaccharides with a relatively low cost [Zhang et al., 1999]. Several plants, insects and microorganisms have chitinolytic enzyme systems that are capable of degrading chitin/chitosan [Azarkan et al., 1997; Patil et al., 2002]. For example, Serratia marcescens produced at least three chitinases with complementary activities [Suzuki et al., 2002]. Degradation of chitin results in a range of chitooligosaccharides with different degree of polymerization. In the past decades, several biotechnological approaches have been taken to prepare industrially chitooligosaccharides. Biodegradation of chitin/chitosan was carried out in batch reactors in the early period of enzymatic production of chitooligosaccharides [Izume and Ohtakara, 1987; Jeon and Kim, 2000; Varum et al. 1996]. However this batch method has some disadvantages: (1) the high cost associated with large quantities of expensive enzymes; (2) low yields due to the limited ability to control the degree of polymer; (3) mixture of chitooligosaccharide with a broad molecular weight range; (4) large quantities of expensive chitosanase used. Especially, the high cost associated with hydrolytic enzymes demotes the application of enzymatic methods. To reduce production cost, it is recommended to reuse hydrolytic enzymes instead of a single use in batch reactors. Enzyme immobilization method was introduced to degrade chitin/chitosan for overcoming the high process cost problems caused by the single used enzyme in batch reactors [Jeon et al. 1998]. In this system the immobilized chitosanase exhibited the highest enzymatic activity. It was found that the effective production of target chitooligosacchrides depended greatly on surface enzyme density, supported particle size, aggregation speed, and initial substrate concentration [Kuroiwa et al., 2002, 2003]. However the poor affinity of immobilized enzymes to chitin/chitosan substrates than that of free enzymes limited this method for hydrolysis of chitin/chitosan effectively. Such problems that the immobilization column reactors faced can be overcome by using ultrafiltration (UF) membrane enzymatic reactor system [Jeon and Kim, 2000]. The enzymatic production of chitooligosaccharides with relatively a high degree of polymerization was processed in this system. The chitin/chitosan-hydrolyzing enzyme added in the reaction vessel was recycled, and the produced oligosaccharides were separated from the substrates and the enzymes in the process of chitosan degradation. The production of reducing sugar progressively increased with arising permeation rate, and the composition of the oligosaccharides was dependent on the permeation rate. Thus permeation rate was a key factor for the control of oligosaccharide production. The molecular weight distribution of the hydrolysate could be controlled within limits by the appropriate membrane used. Large quantities of pentamers and hexamers but no monomers were obtained with the UF membrane reactor system, suggesting this system enable effective production of relatively larger oligosaccharides [Jeon et al. 2001]. Oligosaccharides obtained using this reactor system showed antibacterial activity [Jeon and Kim, 2000; Kittur et al., 2003]. This reactor system could hydrolyze at least 11 batches of substrates for the same amount of enzymes used in the batch reactor. However, UF membrane method did not allow continuous production of chitooligosaccharides because of the increased transmembrane pressure during the reaction. It seems to occur due to the high viscosity of chitosan. Therefore reducing chitosan viscosity prior to treatment in the membrane system may allow membrane fouling to be all eviated.

14

Xianzhen Li and Xiaoyi Chen

Considering the membrane fouling occurred due to the high viscosity of chitosan, researchers have been performed to develop new strategies to degrade chitin/chitosan. As schematically shown in Figure 4, a dual reactor system composed of an UF membrane reactor and a column reactor packed with an immobilized enzyme was developed. With this combination of an immobilized enzyme reactor and the UF membrane reactor, continuous production of chitooligosaccharides was feasible [Jeon & Kim, 2000]. Firstly, chitosan was partially degraded by the immobilized enzyme prepacked in the column reactor, and the hydrolysates with low viscosity were immediately supplied to a substrate feed tank of an UF membrane reactor. Secondly, the partial fragments of chitin/chitosan from immobilization reactor were continuously added to the UF membrane reactor system for the enzymatic degradation to produce chitooligosaccharide. The different molecular weight cutoffs of UF membranes can be used in the system to obtain different molecular weight distribution. This method ensures a greater productivity per unit of enzyme, ability to control molecular weight distribution, and more efficient continuous production process. Kuroiwa et al. [2003] has determined the optimum conditions for continuous production of pentamers and hexamers of chitooligosaccharides using a dual reactor. Under the optimum conditions, continuous production of pentamers and hexamers was achieved for a month without significant decrease in products.

Figure 4. Schematic diagram of the dual reactor system developed for continuous production of chitooligosaccharides. Adapted from Jeon and Kim [2000].

In the enzymatic hydrolysis of chitin/chitosan, the molecular size of the final products is very important because the functional properties of chitin/chitosan and their oligosaccharides are greatly dependent on their molecular weights [Hirano and Nagao 1989]. In enzymatic hydrolysis of chitosan using a batch reactor, chitosan may be cut randomly by endo-type enzymes and then higher oligosaccharides produced may be immediately hydrolyzed by the enzyme in the solution, leading to the preparation of lower oligosaccharides. In a continuous

Biodegradation of Polysaccharide Sourced from Virulence Factor…

15

UF membrane reactor, however, smaller oligosaccharides than the cutoff of the used membrane were permeated and separated from the enzyme during reaction. This probably prevented the oligosaccharides proceeding to further hydrolysis [Izume and Ohtakara, 1987]. For these reasons, the dual reactor system has been mainly used in the enzymatic process for hydrolysis of polysaccharide [Jeon & Kim, 2000].

3) Biodegradation of xyloglucan Xyloglucan is a major structural carbohydrate of hemicellulose occurred in the primary walls of many land plants [Somerville et al., 2004]. It has a cellulose-like backbone of β-1,4linked D-Glcp residues, which is substituted in a regular pattern with xylosyl residues plus other sugars that vary depending on the plant species. Xyloglucan forms a structural network by interaction with cellulose microfibrils via hydrogen bonds, which plays a key role in cell wall integrity [Fry et al., 1993b; Pauly et al., 1999]. Xyloglucan being substituted mainly occurs at C-6 with α-1,6 D-xylosyl (Xylp) residues. Other saccharides, frequently β-Dgalactose (D-Galp) and α-1,2 L-fucose (l-Fucp) residues, may be attached to it [Carpita and McCann, 2000]. In order to conventionally write [Fry et al., 1993b], the unsubstituted D-Glcp residue is named G, α-D-Xylp-(1,6)-β-D-Glcp segment is named X, β-D-Galp-(1,2)-α-D-Xylp-(1,6)-βd-Glcp is L and a-L-Fucp-(1,2)-β-d-Galp(1,2)-α-D-Xylp-(1,6)- β-D-Glcp is F (shown in Figure 5). Two general types of xyloglucans are XXXG-type having a repeating unit with one unsubstituted D-Glcp residue and XXGG-type having two unsubstituted D-Glcp residues [Vincken et al., 1997b]. HO HO OH

OH O

O O

HO OH x OH HO H

O

O

O

O O

O

HO

OH

OH

HO O

O

OH

O HO

O

OH O

O O

OH OH

RO O

O OH OH

OH HO

OH y

OH OH

Figure 5. General structure of xyloglucan including XXXG (x=0, y=0, R=H), XLXG (x=1, y=0, R=H), XXLG (x=0, y=1, R=H), and XLLG (x=1, y=1, R=H). XXFG (x=0, y=1, R=a-1,2 L-Fuc.

n

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Xianzhen Li and Xiaoyi Chen

Xyloglucan-degrading enzymes have attracted the attention of researchers during the last decade. Some cellulases (endo-1,4-β-glucanases) are able to hydrolyze the xyloglucan backbone at unsubstituted D-Glcp [Vincken et al., 1997a]. Two of seven cellulases from Chrysosporium lucknowense possess a notable activity against xyloglucan [Bukhtojarov et al., 2004]. For a long time, three xyloglucan-degrading endoglucanases from plants have been the only examples of specific xyloglucanases acting on the polymer backbone [Edwards et al., 1986; Matsumoto et al., 1997]. A few specific xyloglucanases from fungi have been discovered only recently [Hasper et al., 2002; Yaoi and Mitsuishi, 2004]. Three xyloglucanases isolated from Aspergillus japonicus, Chrysosporium lucknowense and Trichoderma reesei have high specific activity toward tamarind xyloglucan, whereas the activity against carboxylmethyl cellulose and barley β-glucan is absent or very low. All enzymes produced XXXG, XXLG/XLXG and XLLG oligosaccharides as the end products of xyloglucan hydrolysis. Aspergillus japonicus xyloglucanase displayed an endo-type of attack on the polymeric substrate, while the action mode of two other xyloglucanases was similar to the exo-type [Grishutin et al., 2004]. The recent enzymatic characterization of the endoxyloglucanase, Xgh74A, showed that the enzyme hydrolyzed the glycosidic bond of the unbranched glucosyl residues in xyloglucan, to yield XXXG, XLXG, and XLLG oligosaccharides [Zverlov et al., 2005]. Two xyloglucan-specific endo-β-1,4-glucanases were isolated from the Gram-positive bacterium Paenibacillus sp. strain KM21, one of which is a typical endo-type enzyme that randomly cleaves the xyloglucan main chain, while the other of which has dual endo- and exo-mode activities or processive endo-mode activity [Yaoi et al., 2005].

4) Biodegradation of pectin Pectin is a family of complex and highly heterogeneous polysaccharides that contributes to the structure of plant tissues as a component of the middle lamella and primary cell walls [Daas et al., 1999，2001]. Pectin is built of a backbone mainly consisting of β-1,4-linked αD-galacturonic acid (GalUA) residues, with various degrees of methyl esterification of the carboxyl groups on α-D-galacturonic acid residues. The main chain of the polymer also contains rhamnose residues that can be highly substituted by arabinose and galactose side chains. Phytopathogenic fungi produce extracellular enzymes, which can degrade the cell wall components of plants. These fungi not only degrade cell wall polymers to use their sugars as an important nutrient source, but also digest the cell walls to aid in penetrating cells and spreading through the plant tissue. In this rotting process oligomers can be produced by degradation of pectic carbohydrates in cell walls, for example, pectin derived oligosaccharides were generated when cell wall pectins were digested in fruit as they were ripen or they were degraded by pathogen enzymes during tissue colonization [Melotto et al., 1994; Tonukari et al., 2000; Wanjiru et al., 2002]. It is thought that these oligosaccharides are generally important factors for regulating fruit responses to infection by pathogens [OlanoMartin et al., 2002; Ridley et al., 2001]. Degradation of the pectin polymer occurs via a set of pectinolytic enzymes, which can roughly be divided into two groups. One is pectin esterase removing methoxyl groups from

Biodegradation of Polysaccharide Sourced from Virulence Factor…

17

pectin. The other is depolymerase being classified further as lyases (β-elimination) and hydrolases, which degrade the backbone chains (Figure 6) [Sakai et al., 1993]. All hydrolases involved in the degradation of pectin are classified as members of family 28 of the glycoside hydrolases, including the endopolygalacturonases, exopolygalacturonases and rhamnogalacturonases [Markovič and Janeček, 2001]. Polygalacturonate (pectate) lyases are specific for unmethylated pectate, although they can be active on pectin with a low degree of methyl esterification [Tardy et al., 1997]. For example, Pectate lyase C from Bacillus substilis can depolymerize polygalacturonate and pectin of methyl esterification degree from 22% to 89%, but exhibits maximum activity on 22% esterified citrus pectin [Soriano et al., 2006]. Pectate lyases are widely distributed among plant pathogens, where they play an important role as virulence factors [Herron et al., 2000]. They have also been found in saprophytic microorganisms, including members of the genus Bacillus and in some thermophilic bacteria [Berensmeier et al., 2004; Hatada et al., 2000; Kluskens et al., 2003]. Pectin lyases cleave the 1,4-α-linkages in pectin molecules and are mainly synthesized by fungal species, whereas few bacteria produce this type of enzyme. The pectin lyases have a specificity on pectin molecules, for example, an exo-acting pectin lyase from an Aspergillus sp. acted on highly esterified pectin but not on polygalacturonic acid [Delgado et al., 1992]. In the culture of anaerobe Bacteroides thetaiotamicron, there were both polygalacturonate hydrolysase and polygalacturonate lyase activities to be detected, which have been identified as membrane-associated enzymes [McCarthy et al., 1985]. Phytopathogenic Pseudomonas spp. causing soft rots of plant secreted a constitutive pectate lyase when grown in calcium-containing media [Liao, 1991]. R1 O

R1

R1 COX

HO

OH

COX

COX OH

O

HO

OH

OH COX HO

OH

Pectin lyase Exo-polygalacturonate lyase Pectate lyase

O

Pectinase Exo-polygalacturonase

O

O

O

O

COX

COX

OH

O

HO

O

R2

R2

Pectin methyl esterase

R1

OH OH

O

HO

O

HO

O OH

O

R2

O -

COO HO

O OH

O HO

COO

-

O OH

O R2

Figure 6. Pectin degradation by pectiolytic enzymes (X is predominantly O- for pectate and O-Me for pectin, R1 and R2 are polygalacturonate).

In studying the pectin-degrading enzymes produced by plant pathogens, such as Erwinia chrysanthemi, the hydrolytic action and production of pectinases and polygalacturonate lyase

18

Xianzhen Li and Xiaoyi Chen

have provided a mechanism for the bacteria to degrade the plant cell wall and effect on pathogenicity, and also bacteria utilize such breakdown products for bacterial growth. It was found that pectate lyases produced by soft-rotting bacterium Erwinia spp. play a significant role in pathogenic infection [Reverchon and Baudouy, 1987]. The enzymes are induced by breakdown products of polygalacturonic acid such as 2-keto-3-deoxygluconate, and their production is repressed by glucose. What is perhaps surprising is that the bacteria should synthesize five pectate lyases, apparently in addition to an oligogalacturonate lyase active against the unsaturated digalacturonate. As well as playing a very important role in the plant infection process, the lyases enable the bacteria to grow on pectin as the sole carbon and energy source. Oligogalacturonides with degrees of polymerization between 10 and 20 have been shown to have important regulatory activity for plant defense mechanisms [Cervone et al., 1989].

5) Biodegradation of xylan Xylan, a β -1,4 linked polymer of D-xylose with D-glucuronic acid or L-arabinose substituents, occurs in almost all parts of plant cell walls, which is the one main component of plant hemicelluloses. Production of xylooligosaccharides from plant hemicelluloses generally includes extraction of xylans and subsequent enzymatic hydrolysis of the extracted xylans [Yuan et al., 2004]. Xylan can be depolymerized by the enzymatic degradation of xylanases. The enzymatic degradation of plant cell wall xylan requires the concerted action of a dedicated enzymatic consortium (diverse enzymatic syndicate) due to the constituent groups in xylan acting as a limited factor in efficient hydrolysis. Among these enzymes there are endo-1,4-β-xylanses and β-xylosidases that cleave the backbone chain, and β-Larabinofuranosidase, xylan esterases, and β-D-glucuronidase that cleave the side chain of xylan (Figure 7) [Coughlan and Hazlewood, 1993]. Esterases hydrolyze the O-acetyl substituents, primarily at the O-2 position of the xylan backbone, which include two distinct ferulae and acetylxylan esterases catalyzing to remove the ferulate groups linked via arabinosides to the xylan backbone and to deacetylate the O-3 and (primarily) O-2 positions of the xylan backbone in acetylxylan [Taylor et al., 2006]. Several bacterial and fungal species produce the full complement of enzymes necessary to utilize xylan as a carbon source [Uffen, 1997]. Xylanase activity has been detected in some members of the genus Paenibacillus, Bacillus, Xylanibacter, Streptomyces [Jiang et al., 1994; John et al., 2006; Lee et al., 2000; Morales et al., 1995; Ueki et al., 2006; Velázquez et al., 2004].

Biodegradation of Polysaccharide Sourced from Virulence Factor… -

O

O

α-Glucuronidase

O

MeO HO

Acetyl xylan esterase H3C Xylanase

OH O

HO O

O

O

O MeO HO

O

O

O

19

OH O

OH

HO O

O O O O

n

α-L-Arabinofuranosidase OH

Ferulate esterase

Figure 7. Schematic diagram of xylan degradation.

Endoxylanases are classified into two groups, family 10 or 11, based on hydrophobic cluster analysis and amino acid sequence homologies [Henrissat, 1991]. Compared to endoxylanases of the family 11, those belonging to the family 10 show better capability of cleaving glycosidic linkages in the xylan main chain closer to the substituents, such as MeGlcA and acetic acid. The thermophilic fungi, Thermoascus aurantiacus and Sporotrichum thermophile, produce high activities of highly thermostable xylanases under solid state and submerged culture respectively. A high molecular xylanase (XYL I) with catalytic properties similar to those belonging to the family 10 was purified and characterized by the culture filtrates of Thermoascus aurantiacus, and a low molecular weight xylanase (XYL A) with catalytic properties similar to those belonging to the family 11 was isolated from the culture filtrates of Sporotrichum thermophil [Kalogeris et al., 2001]. When treatment with a Thermoascus aurantiacus family 10 or a Sporotrichum thermophile family 11 endoxylanases, acidic oligosaccharides can be degraded from birch wood xylan. The main difference between the products liberated by xylanases of the family 10 and that from the family 11 concerned the length of the products containing 4-O-methyl-D-glucuronic acid. The xylanase from Thermoascus aurantiacus liberated an aldotetrauronic acid as the shortest acidic fragment from glucuronoxylan in contrast with the enzyme from Sporotrichum thermophile, which liberated an aldopentauronic acid. It was found that xylooligosaccharides could be typically produced by enzymatic degradation of lignocellulosic materials, having xylan as major hemicellulose component [Parajó et al., 2004]. However, such degradation was concerned with the substituent in polysaccharides. The substituted xylooligosaccharides, arabinoxylooligosaccharides, can be formed from wheat flour arabinoxylans. Arabinoxylans consist a backbone of (1,4)- β-linked D-xylopyranosyl units with substitutions of α-L-arabinofuranosyl units at position C-(O)-2 and/or C-(O)-3 [Courtin and Delcour, 2001; Gruppen et al., 1992]. Endoxylanases attacked the arabinoxylans backbone internally, depending on the arabinose substitution pattern [Swennen et al., 2005]. They degraded arabinoxylans yielding series of oligomeric and polymeric arabinoxylans with different DP [Courtin and Delcour, 2001; Maes et al., 2004].

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Xianzhen Li and Xiaoyi Chen

Debeire et al [1990] purified an extracellular xylanase from a hydrolytic thermophilic anaerobe, Clostridium thermolacticum. They observed that the xylanase was an endo-1,4xylanase and the main hydrolyzed end products of the larchwood xylan were xylobiose and xylotriose. Pellerin et al [1991] employed this xylanase for production of xylooligosaccharides from xylan derived from corncob meal by alkaline extraction. However, a long hydrolysis time of 48 h was required with 100 000 U of the xylanase. The substrate specificity of the xyloglucanase Cel74A from Hypocrea jecorina (Trichoderma reesei) was examined using several polysaccharides and oligosaccharides. The results revealed that xyloglucan chains were hydrolyzed at substituted glucose residues, in contrast to the action of all known xyloglucan endoglucanases [Desmet et al., 2007]. It has been proved that the xylooligosaccharides with DP less than 5 failed to induce the xylanase activity for plant defense [Miyazaki et al., 2005].

6) Biodegradation of carrageenan Carrageenans are sulfated galactans occurred in the cell wall of marine red seaweeds. They are major components of the matrix involved in the building up of the cell-wall architecture and mediate cell-cell recognition in host-pathogen interactions [Kloareg et al., 2001]. Carrageenan is a generic name for a family of natural, water-soluble, sulfated galactans [De Ruiter and Rudolph, 1997]. This large family of anionic polymers share the same backbone structure, which consists of a linear chain of alternating α-(1,3)-linked β-Dgalactopyranose (Galp) and β-(1,4)-linked α-D-galactopyranose. In these units, various hydroxyl groups may be substituted by ester sulfate [Knutsen et al., 1994, 1995]. Usually a 3,6-anhydro-α-D-galactopyranose (AnGalp) in place of α-D-galactopyranose may contains in linear chain. Carrageenan can be classified according to the number and the position of sulfated ester. For example, the three most industrially exploited carrageenans, namely kappa(κ, A-G4S), iota- (ι, A2S-G4S) and lambda- (λ, D2S6S-G2S) carrageenans, are distinguished by the presence of one, two and three ester sulfate groups each repeating disaccharidic unit respectively (Figure 8). However, carrageenans have very heterogeneous chemical structures, depending on the algal sources, the life stages and the extraction procedures of the polysaccharides. It was reported that carrageenan could be degraded by using an active oxygen species or organic acid solution [Yamada et al. 2000; Yu et al. 2002]. However, the degradation degree of carrageenan with these chemical methods was controlled difficultly, and usually very small molecules even its constituent monomer was produced, in which the products were a complex mixture. Therefore it was thought that the most popular technical method should be the enzymatic hydrolysis of carrageenan. Several genera of marine bacteria have been found to produce carrageenase, such as Vibrio, Alteromonas, Pseudomonas carrageenovora and Cytophaga [Araki et al. 1999; Michel et al. 2000, 2001; Potin et al. 1991].

Biodegradation of Polysaccharide Sourced from Virulence Factor…

a

OSO3CH2OH O

O O

O

O

OH

n

OH

b

OSO3CH2OH O

O O

O OH

c

OH

O OSO3-

OSO3-

CH2OH O

n

O O OSO3-

O OSO3-

21

HO

n

Figure 8. Disaccharide repeating units of carrageenans. a, k-Carrageenan (β-D-galactopyranosyl-4sulfate(1,4)-O-3,6-anhydro-α-D-galactopyranosyl); b, ι-Carrageenan (β-D-galactopyranosyl-4-sulfate(1,4)O-3,6-anhydro-α-D-galactopyranosyl-2-sulfate); c, λ-Carrageenan (β-D-galactopyranosyl-2-sulfate(1,4)-O-aD-galactopyranosyl-2,6-disulfate).

To prepare sulfated oligosaccharides with special biological activities, a marine bacterium Cytophaga strain MCA-2 was isolated, which produced a special extracellular carrageenase in the presence of crude κ-carrageenan [Mou et al. 2002]. This κ-carrageenase can cleave specifically the internal α-(1,4) linkages of κ-carrageenans to yield oligosaccharides of the neocarrabiose series [Barbeyron et al., 2000; Michel et al., 2001]. Enzymatic products of κ-carrageenans have been shown to elicit biological activities in plants [Bouarab et al., 1999]. Depolymerization of κ-carrageenan was performed also using other carrageen-degrading enzymes for oligosaccharides production [Ekeberg et al., 2001; Knutsen et al., 2001]. In order to prepare the oligosaccharides with different molecular weights and sulfur contents, carrageenan was first fragmented by enzymatic hydrolysis, and then followed by sulfonation with formamide-chlorosulfonic acid. Semi-preparative chromatography based on a strong

22

Xianzhen Li and Xiaoyi Chen

anion-exchange (SAX)-HPLC column has been used to separate such oligosaccharides [Yu et al., 2002]. Carrageenan derived oligosaccharides have been shown to have biological activities such as being an disintegrant and anti-viral [Shi et al. 2000]. In general, the biological activities of sulfated oligosaccharides are close related with the molecular weight, the carbohydrate structure and the content and linking position of sulfur groups [Liu et al., 2000].

7) Biodegradation of xanthan Xanthan is an extracellular heteropolysaccharide produced by the phytopathogenic bacterium Xanthomonas campestris pv. campestris. Xanthan molecule consists of a main cellulosic backbone with linear trisaccharide side chains, each of which is composed of a mannosyl (β-1,4)-glucuronyl-(β-1,2)-mannose sequence attached at the C-3 position on alternate glucosyl residues through α-1,3 linkages (Figure 9) [García-Ochoa et al. 2000]. The molecular weight distribution ranges from 2 × 106 to 20 × 106 Da. The internal mannosyl residues of the side chain are mostly acetylated at O-6 position, and approximately 50% of the terminal mannosyl residues may be substituted with a pyruvate ketal at C-4 and C-6 positions. The levels of pyruvate and acetyl substitution are varied depending on growth conditions and bacterial strains [García-Ochoa et al. 2000]. It has been proved that exopolysaccharide xanthan is the virulence factor of Xanthomonas campestris pv. campestris causing block rot process in cruciferous plants, which leads to a significant loss of crop in the worldwide [Chou et al., 1997; Ishikawa et al., 2004]. It has been believed that the defense response in plant cells mediated by elicitor molecules could help plant resist disease [Fry et al. 1993b]. Like some elicitor-active oligosaccharides produced artificially by the enzymatic fragmentation of glucan, chitosan and pectin [Boudart et al. 1998; Fry et al. 1993b; Küpper et al. 2001], xanthan molecule has the similar structural features to these polysaccharides such as the side chain and charged residues [García-Ochoa et al. 2000], presumably xanthan degradation products have elicitor activity. In fact, some xanthooligosaccharides derived from xanthan are elicitor-active to induce the accumulation of phytoalexin in the soybean cotyledon [Liu et al., 2005]. The xanthan degradation products have also been shown to directly inhibit Xanthomonas. campestris pv. campestris, but none of the other organisms was tested in vitro [Qian et al., 2006]. It is interesting that xanthan produced by Xanthomonas campestris pv. campestris was involved in the pathogenicity [Ishikawa et al., 2004], but its partial hydrolysates could inhibit the cell growth of Xanthomonas campestris pv. campestris [Qian et al., 2006]. Therefore the xanthan degradation products were potential protecting agent for the management of diseases caused by Xanthomonas campestris pv. campestris [Liu et al., 2005; Qian et al., 2006].

Biodegradation of Polysaccharide Sourced from Virulence Factor… CH2OH O HO

OH CH2OAc

HO

O Man OH

Glc

CH2OH O Glc HO OH

CH3

HO

O HO

O Man OH

CH2OH O O

OH CH2OAc HO Man HO

OH

COOH HO

OH

β -D-Glucanase

Glc

O

OH

CH2OAc O HO Man OH HO

CH2OH O COOH

n

O O

β -D-Glucanase

O

O OH

O

COOH O GlcA O HO OH

O

O

O

HO Man HO

COOH CH3

CH2OH O

O

Glc

Xanthal lyase

23

O

Glc

O OH

n

O O

CH2OH O Glc OH

α -D-Manosidase

CH2OAc O O HO Man OH HO

O

GlcA

OH COOH

HO

O OH

GlcA OH

OH

β -D-Glucosidase CH2OH O HO Glc HO

O OH CH2OAc HO Man HO

COOH HO

O

CH2OH O Glc OH

CH2OH O OH

Glc

O O

CH2OAc HO Man HO

O

COOH

GlcA OH

OH OH

HO

O

GlcA OH

O O

CH2OH O HO Glc HO

O

OH

Unsaturated glucuronyl hydrolase

Figure 9. Xanthan depolymerization pathway in Bacillus sp. strain GL1 [Nankai et al., 1999].

Xanthan is a highly stable polysaccharide that can be completely degraded into its smallest constituents by a few microorganisms [Ahlgren, 1993; Cadmus et al., 1982; Cheetham and Mashimba, 1991; Christensen et al. 1996; Nankai et al., 1999; Sutherland, 1984], although in the unordered conformation it can be degraded partially by cellulase [Rinaudo and Milas, 1980]. Cadmus et al. [1982] described the isolation of a salt-tolerant bacterium Bacillus sp. K11 capable of eliminating side chains from xanthan. The xanthandegrading bacterium Paenibacillus alginolyticus degrades approximately 28% of the xanthan molecule and appears to leave the backbone intact [Ruijssenaars et al. 1999b]. Bacillus sp. GL1 was found to utilize xanthan for its growth and thereby produced xanthan lyase [Hashimoto et al. 1998a]. Although some microbial mixed cultures have been found to assimilate xanthan, an enzymatic route for the complete depolymerization of a xanthan in Bacillus sp. strain GL1 was elucidated in 1999 by analyzing the structures of xanthan depolymerization products [Nankai et al., 1999]. As shown in Figure 9, the glycosidic bond between pyruvylated mannosyl and glucuronyl residues in xanthan side chains was first attacked by extracellular

24

Xianzhen Li and Xiaoyi Chen

xanthan lyase to remove the pyruvylated mannose. And then, this modified xanthan was depolymerized to a tetrasaccharide by extracellular β-D-glucanase, without the terminal mannosyl residue of the side chain in a repeating unit of xanthan. After that, the tetrasaccharide was incorporated into cells and one glucose residue was cleaved from nonreducing end by β-D-glucosidase. The produced trisaccharide (unsaturated glucuronylacetylated mannosyl-glucose) was degraded successively to unsaturated glucuronic acid and a disaccharide (mannosyl-glucose) by unsaturated glucuronyl hydrolase. At last, the disaccharide was hydrolyzed to the constituent monosaccharides (mannose and glucose) by α-D-mannosidase. According to the enzymatic route of xanthan, there are five xanthan-degrading enzymes involved in xanthan degradation. Xanthanase (β-D-glucanase) catalyses the hydrolysis of the xanthan backbone. A few xanthanases have been identified, some of which were categorized as cellulase family members [Sutherland, 1987]. It was found that some cellulase showed partially hydrolytic activity on xanthan [Rinaudo and Milas, 1980]. The endoglucanases probably acted in conjunction with xanthan lyase and showed a higher activity on xanthanderived oligosaccharides than on intact xanthan [Ahlgren, 1991; Hashimoto et al. 1998b; Sutherland, 1987]. However, the enzyme from Paenibacillus alginolyticus XL-1 was active only on intact xanthan and was not found to be associated with endoglucanases [Ruijssenaars et al., 1999]. Xanthan lyase is an enzyme first attacking xanthan, which removes the terminal mannosyl residue via β-elimination and yields a free mannose and an unsaturated glucuronyl terminal in the side chain of xanthan [Ahlgren, 1991; Sutherland, 1987]. Xanthan lyase is usually used as a xanthan-modifying enzyme for studying structure-function relationships and producing modified xanthan. There are two types found in xanthan lyase, non-specific and pyruvated mannose-specific xanthan lyases [Ahlgren, 1991; Hashimoto et al., 1998b; Ruijssenaars et al., 1999; Sutherland, 1987]. It was clear that the acetyl group was not required for lyase activity, since the enzyme was active on xanthans that were originally low in acetyl substituents as well as on chemically deacetylated xanthan [Ruijssenaars et al., 1999]. The cloning and sequencing of xalA gene was first described in Paenibacillus alginolyticus XL-1, encoding pyruvated mannose-specific xanthan lyase. The mature enzyme could be expressed functionally in Escherichia coli showing no activity on depyruvated xanthan like the native enzyme [Ruijssenaars et al., 2000]. Soon after, another gene for the xanthan lyase was cloned from Bacillus sp. strain GL1 in which the lyase was induced by xanthan [Hashimoto et al., 2001]. Pyruvated mannose-specific xanthan lyases have been purified from a salt-tolerant mixed culture and Bacillus sp. [Ahlgren, 1991; Hashimoto et al., 1998b]. Mixed or pure bacterial cultures grown on xanthan generally produced a mix of xanthandegrading enzymes [Hashimoto et al., 1998b; Ruijssenaars et al., 1999]. Cadmus et al. [1982] described the isolation of a xanthan-degrading bacterium. The xanthanase occurred in culture was a mixture of the enzymes attacking all of the side chain linkages in the xanthan molecules, including the xanthan lyase, α-D-mannosidase, and unsaturated glucuronyl hydrolase. They found no depolymerase activity in their cultures because the β-1,4-linked glucan backbone remained intact. Also the enzymes excreted by Paenibacillus alginolyticus XL-1 only removed residues from the xanthan side chains, whereas long stretches of the β1,4-glucan remained intact [Ruijssenaars et al., 1999].

Biodegradation of Polysaccharide Sourced from Virulence Factor…

25

Recently a xanthooligosaccharide has been produced by degradation of a virulence factor of xanthan [Liu et al., 2005], which in turn inhibited the cell growth of the xanthan-producing phytopathogen Xanthomonas campestris pv. campestris [Qian et al., 2006]. Such inhibitory activity was greatly dependent on their hydrolytic degree, which was critical for microorganism inhibition [Jeon et al., 2001]. It was suggested that the degraded xanthan was potentially valuable biological control agent, because Xanthomonas campestris pv. campestris is one of the most important phytopathogen causing crucifer black rot, and the xanthan is its important virulence determinant.

8) Biodegradation of alginate Alginate is a linear anionic copolymer composed of (1,4) linked β-D-mannuronic acid (M) and its C-5 epimer α-L-guluronic acid (G). As shown in Figure 10, they consist of the alternation of homopolymeric blocks of poly-β-1,4-D-mannuronic acid (referenced to MM blocks, a), of homopolymeric blocks of poly-α-1,4-L-guluronic acid (GG blocks, b), and of heteropolymeric blocks with random arrangements of both monomers (MG blocks, c). The proportion and sequence of the block structures vary greatly in alginate molecules, depending on alginates derived from the seaweed sources [Gacesa, 1988]. Like other polysaccharide derived from fungi and plant [Sharp et al., 1984a; Shibuya and Minami, 2001], alginate is also the cell wall constituent of brown seaweeds and as an exopolysaccharide produced by pathogen during infection process [Gacesa, 1988]. One of the critic feature of alginate is having many negative charges occurred in alginate molecules. It has been proved that alginate produced by brown seaweeds is not acetylated [Onsøyen, 1996]. The acetylated form of alginate is synthesized by certain bacteria, such as mucoid cells of Pseudomonas aeruginosa and Azotobacter vinelandii [Cote and Krull 1988; Pier, 1998]. Alginate has also been found to function as a major virulence factor of several pathogenic bacteria during the infectious process such as Pseudomonas aeruginosa, Pseudomonas syringae [Keith et al., 2003; Peñaloza-Vázquez et al., 2004; Pier, 1998]. Marine alga alginate is widely used in food, cosmetics and pharmaceutical industries owning to its gelling ability, stabilizing properties and high viscosity [Ci et al., 1999]. However, the increasing attention has been paid to alginate oligomers produced by alginate lyases, because it has been revealed that alginate hydrolysates and their derivatives exhibit many important bioactivities, such as stimulating the growth of plant root or Bifidobacterium [Akiyama et al., 1992], and causing cytotoxic cytokine production in human mononuclear cells [Iwamoto et al., 2003]. Alginate lyase also has attracted public attention recently because the enzymatic degradation of alginate expands the potential application of this polysaccharide [Hu et al., 2004a, 2004b]. Depolymerized alginates with low molecular weight act like oligosaccharins in their ability to regulate physiological process in plants [Hu et al., 2004a].

26

Xianzhen Li and Xiaoyi Chen

a

-

COO

O HO

M

b

HO

O

OH

O

COOO

OH - O COO

COO-

OH

G

OH

O

O

O

OH

GO

O HO

COO- O OH

M

OH COO-

O

OH

GO O

O

O

G

OH COO-

COO-

OH

c

O

OH

OH

O

G

O HO

M

COO- O M OH

O HO

M

O OH

O

OH COO-

Figure 10. Chemical structure of alginates from brown algae.

Alginates can be degraded enzymatically by alginase, however there are no other alginate hydrolyase being reported except alginate lyase, which catalyze the β-elimination of the 4-Olinked glycosidic bond to form a unsaturated double bond between C-4 and C-5. In this lytic action molecule containing 4-deoxy-L-erythro-hex-4-enepyranosyluronate at the nonreducing end was generated [Gacesa et al., 1989]. As already indicated, alginate lyases are of widespread occurrence, being found in marine gastropods, bacteriophage, various marine microorganisms and some soil. Up to date, alginate lyase has been found to occur in many microorganism, such as Alginovibrio aquatilis, Bacillus circulans, Dendryphiella salina, Enterobacter cloacae, Klebsiella aerogenes, Klebsiella pneumoniae, Pseudomonas aeruginosa, Pseudomonas alginovora, etc. [Boyd and Turvey, 1977; Boyen et al., 1990; Caswell et al., 1989; Hansen et al., 1984; Linker and Evans, 1984; Nibu et al., 1995; Shimokawa et al., 1997; Stevens and Levin, 1977]. Not all of the reported alginate lyases have been extensively studied, but those which have been characterized seem to be capable of degrading both bacterial and/or algal alginates and partially hydrolyzed polyuronic acid substrates [Davidson et al., 1976; Dunne and Buckmire, 1985; Hansen et al., 1984]. We recently isolated a bacterium of Flavobacterium sp. LXA as a high alginate-degrading enzyme producer from decayed seaweed. Alginate could be decomposed controllably by alginase from the strain LXA to form alginooligosaccharide. It has been found that some of those oligosaccharides are elicitor-active on plants, inducing the accumulation of phytoalexin in the soybean cotyledon (unpublicated results). The alginate lyases have an absolute specificity for either a D-mannuronate or an Lguluronate residue at the non-reducing side of the bond to be cleaved, although usually there is no restriction on the uronic acids present at the reducing side [Boyd and Turvey, 1977]. Moreover, most of the alginate lyases are randomly endolytic enzymes, whereas some with

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exo-activity have also been reported [Doubet and Quantrano 1984]. To date, three types of alginate lyases have been identified [Østgaard, 1993], the specificities of which are defined in terms of the preference for polymannorate block or polyguluronate block. It was difficult to determine which linkage bond of the M-G, the G-M or mixed random PolyMG being degraded. The alginate lyases found in Photobacterium sp. and Pseudomonas aeruginosa are representatives of poly M lyase [Linker and Evans, 1984; Romeo and Preston, 1986], and those found in Klebsiella aerogenes are representatives of poly G lyase. [Lange et al., 1989] The alginate lyase purified from Azotobacter vinelandii, Alteromonas sp. strain H-4 and Vibrio sp. 510-64 was shown to be capable of degrading both poly M and poly G, which was similar to an enzyme from Bacillus circulans [Ertesvåg et al., 1998; Hu et al., 2006; Sawabe et al., 1997]. The efficiency of degradation depends on the block type [Larsen et al., 1993]. The polymannuronate specific lyases from Dolabella yielded incomplete degradation of substrates, the main products being di-and tri-uronides but no monomeric material [Nisizawa et al., 1968]. It was found that Sphingomonas sp. strain A1 depolymerized alginate by three types of cytoplasmic alginate-depolymerizing enzymes (alginate lyases A1-I, A1-II, and A1-III) [Hashimoto et al., 2000]. Briefly, A1-I was active on acetylated and non-acetylated alginates. A1-II preferred polyG and non-acetylated alginate produced by brown seaweeds. A1-III efficiently liquefied polyM and acetylated alginates produced by mucoid cells of Pseudomonas aeruginosa [Yoon et al., 2000]. These three alginate lyases produced di- and trisaccharides from alginate as major final products [Hashimoto et al., 1998b; Yoon et al., 2000], which implied that the cells of Sphingomonas sp. strain A1 had an additional enzyme responsible for the degradation of alginate oligosaccharides to the constituent monosaccharides. Although several bacterial and algal alginate can be degraded by the enzymes, the specificity of the alginate lyase was not identical to different source of alginate. When compared to other polysaccharide lyases, the mode of action of enzymes degrading alginate is further complicated by the differences in distribution of monosaccharide residues in different substrates. Thus, alginates with the same or very similar mannuronate:guluronate ratios can have very great differences in monosaccharide sequence and in the frequency of adjacent residues. It is true that the choice of enzyme sources and reaction conditions affect the end products. Therefore, the elucidation of the substrate specificities of alginate lyases toward different kinds of oligouronic acids is important for the preparation of desired oligouronic acids by enzymatic degradation of alginate.

5. WELL-KNOWN OLIGOSACCHARIDES AND THEIR STRUCTURAL EFFECT ON BIOLOGICAL ACTIVITIES Higher plants have the ability to initiate various defense reactions, such as hypersensitive responses, production of phytoalexins and antimicrobial proteins, and reinforcement of cell walls when they are infected by various pathogens [Dangl and Jones, 2001]. They can distinguish self and non-self, or detect specific pathogens through the perception of signal molecule elicitors [Kaku et al., 2006]. Fragments of cell surface macromolecules typical of microorganisms such as cell wall polysaccharides often serve as a potential elicitor to induce

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defense reactions. Oligosaccharides as elicitor molecules have been shown to induce defense responses in plant cells [Hahn, 1996]. Some oligosaccharides produced artificially by the enzymatic fragmentation of the constituent of plant or fungi cell wall, as well as the virulence factor of extracellular polysaccharides produced by pathogen, have been demonstrated to be elicitor- or antimicrobe-active [Boudart et al., 1998; Fry et al., 1993b; Liu et al., 2005; Qian et al., 2006]. Their size, polyanionic or polycationic characteristics, and molecular shape are important structural features for the biological activity of oligosaccharides [Aldington et al., 1991].

1) β-glucan-derived oligosaccharides Oligosaccharide elicitors derived from the β-glucans of pathogenic mycetes have been very well characterized, in which a branched hepta-β-glucoside generated from Pseudomonas sojae glucan by partial hydrolysis is the most active elicitor in soybean cells [Sharp et al., 1984a]. Plant cells can recognize the specific features of the hepta-β-glucoside structure, including all three non-reducing terminal glucosyl residues and their spacing along the backbone of the molecule [Cheong et al., 1991]. However, the hepta-β-glucoside did not act as an elicitor in tobacco cells, but a linear β-1,3-linked glucooligosaccharide (laminari oligosaccharides) was elicitor-active on tobacco cells [Klarzynski et al., 2000]. Moreover, it has not been demonstrated any β-glucan fragment to be elicitor-active on monocots up to now. Inui et al. [1997] reported the induction of chitinase and PAL activity in cultured rice cells by laminari hexaose. Recently, Yamaguchi et al. [2000] reported that a reduced pentasaccharide (glucopentaose) derived from the β-glucan of Magnaporthe grisea (Pyricularia oryzae) can induce phytoalexin biosynthesis as elicitor in rice cells at 10 nM concentrations. Whereas its structure is quite different compared with hepta-β-glucoside. Hepta-β-glucoside is a 1,6-linked β-glucooligosaccharide with branches at the C-3 of two 6linked glucosyl residues, but glucopentaose is a 1,3-linked β-glucooligosaccharide with branch at the C-6 of a 3-linked glucosyl residue. Kobayashi et al. [1993] examined the influence of the terminal group of glucan oligosacxcharides on the induction of phytoalexin biosynthesis in alfalfa by comparing a pyridylaminated hepta-β-1,3–1,6 linear glucan with the non-modified hepta-β-glucoside, and found the latter being far less active than the former. The similar results were obtained in the bean cotyledon and the essential minimal structure for biological activity was shown to be a β-1,3–1,6 triglucoside [Tai et al., 1996a, b]. An elicitor active β-1,3-, 1,6-oligoglucans with DP of between 8 and 17 on tobacco was isolated from Alternaria alternata, in which the 1,6linked and non-reducing terminal residues are essential for the elicitor activity. Its activity was about 1000 times more potent than that of laminarin [Shinya et al., 2006]. Degree of polymerization also is an important attribute for biological activities of oligosaccharides. In general, oligosaccharides able to induce a biological response have a DP higher than 4 [Darvill et al., 1992]. An α-(1,3)-linked D-glucan with an average DP of 23 glucose units inhibited completely local lesion development of potato virus Y on Nicotiana tabacum [Singh et al., 1970]. β-1,3-Linked glucan oligomers with a DP over 4 stimulate the release of chitinase, and DP 6 trigger the expression of PAL activity in rice cells [Inui et al., 1997].

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Three non-reducing terminal glucosyl residues, as well as the distribution of the sidechains along the backbone of the molecule, are essential for the elicitor activity [Cheong et al., 1993, 1991] when using a family of chemically synthesized oligo-β-glucosides ranging in size from hexamer to decamer for elucidation elicitor activity. In contrast, the reducing terminal glucosyl residue of the hepta-β-glucoside elicitor is not essential for activity [Cheong et al., 1991].

2) Chitin and chitosan oligosaccharides Chitin/chitosan oligosaccharides, chitooligosaccharide, can inhibit the growth of some strains of bacteria, fungi, and yeasts [Kendra and Hadwiger, 1984; Roller and Covill, 1999; Sudarshan et al., 1992; Wang, 1992]. The reported microorganisms inhibited by chitooligosaccharides include Bacillus subtilis, Candida sp., Enterobacter sakazakii, Escherichia coli, Lactobacillus sp., Listeria monocytogenes, Micrococcus luteus, Pseudomonas aerginosa, Rhodotorula sp., Salmonella typhimurium, Staphylococcus aureus, Staphylococus epidermidis, Streptococus mutans, Streptococcus faecalis, etc. [Jeon et al., 2001; No et al., 2002; Rhoades & Roller, 2000; Tsai et al., 2000]. The antimicrobial activity of chitooligosaccharides is regulated by the subtle difference in oligosaccharide structure such as degrade of polymerization, charge number and distribution, nature of chemical modification to the molecule, and type of microorganism [Chung et al., 2004; Gerasimenko et al., 2004; Jeon et al., 2001; Muzzarelli, 1996; Park et al., 2004a, 2004b; Tsai et al., 2002]. Uchida et al. [1989] observed that chitooligosaccharides mainly with a DP of 4-6 possessed maximal antimicrobial and antifungal effects, while those mainly with a DP of 3-4 showed no activity. It was found that chitooligosaccharide with average molecular weight of less than 2200 Da was hard to suppress microbial growth, but those with molecular weight of around 5500 Da suppressed the cell growth. In another study, Kendra and Hadwiger [1984] examined the extent to which DP can be reduced before antifungal activity was adversely affected using phytopathogen Fusarium solani. The shortest chitooligosaccharide exhibiting the maximum antifungal activity was the heptamer and then the antimicrobial activity decreased with chain length, where the dimer and trimer were inactive. Antibacterial activity of chitosan oligomer is generally superior to chitin oligomer because chitosan possesses a lot of polycationic amines, facilitating their binding on the bacterial cell surface with the negatively charged residues of macromolecules, and subsequently inhibiting bacterial growth [Young and Kauss, 1983]. The death rate of bacterial cells tends to increase upon the increase in degree of deacetylation of chitooligosaccharides [Tsai et al., 2002]. In most cases, 85-95% deacetylation has shown to be responsible for performing the highest antibacterial activity [Chung et al., 2004]. Also the increase in antibacterial activity of chitooligosaccharides is correlated with the arising molecular weight in the same deacetylation case [Park, et al., 2004b]. The charge distribution of chitooligosaccharides is in conjunction with its antifungal activity [Hirano & Nagao, 1989]. Structural modification of chitooligosaccharides by introducing positively charged groups can improve the antibacterial activity.

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In addition, charge distribution of bacterial cell wall plays a considerable role in antibacterial activities. Chung et al. [2004] studied the cell surface characteristics and revealed a close relationship between hydrophilicity and negative charge distribution of bacterial cell surface. The negative charge distribution on cell surface of Gram-negative bacteria was higher than that of Gram-positive bacteria, leading to a higher hydrophobicity on former cell surface. Moreover, inhibition degree of chitooligosaccharides against tested bacteria fits very well the order of high negatively charged Gram-negative bacteria to less negatively charged Gram-positive bacteria [Wang, 1992].

3) Pectin oligosaccharide Pectin oligosaccharide, oligogalacturonides, derived from pectic polysaccharides of plant cell walls have also been known to act as elicitors to induce biosynthesis of phytoalexins, proteinase inhibitors, and lignification [Bruce and West, 1989; Dixon et al., 1989; Farmer et al., 1991]. In general, oligogalacturonides is the most active only if the number of galacturonic acid residues is not lower than 10 [Cervone et al., 1989]. Whereas Reymond et al. [1995] showed that the ability of oligogalacturonides to induce protein phosphorylation increased with the size of the oligogalacturonides even to a DP more than 20. Oligogalacturonides usually require higher concentrations to show elicitor activity compared to chitin- or glucan-oligomers. Compared with the underived oligogalacturonides, the modified oligogalacturonides on the terminal residues reduce the biological activity. The introduction of a 4,5-unsaturated bond in the non-reducing terminal residues by pectate lyase can lower the most active size of oligogalacturonides from DP 10 to 12 [Spiro et al., 1998]. Although some oligogalacturonides able to stimulate defense response have been characterized, a little is known about signal transduction in these systems. Some evidences supported that these molecules activated defense genes in plants via the octadecanoic pathway, which increases the jasmonic acid level in leaves of tomato plants through an intermediary formation of linolenic acid hydroperoxidase [Doares et al., 1995]. The increase of jasmonic acid levels leads to the transcriptional activation of defense genes [Doares et al., 1995]. Up to date, there has been no reports of elicitor activity of oligogalacturonides in monocot plants, which is identical to the much lower content of pectic polysaccharides occurred in monocot cell walls.

4) Alginate-derived oligosaccharide Most known bioactive oligosaccharides are sourced from the structural components of pathogen or plant cell walls [Fry et al., 1993b]. And an active oligosaccharide has also been degraded from a virulence factor of extracellular polysaccharide produced by phytopathogen [Liu et al., 2005; Qian et al., 2006]. Alginate is both the component of algal cell walls and the virulence factor of some pathogen [Gacesa, 1988; Peñaloza-Vázquez et al., 2004], therefore it should be also elicitor-active on host plants or have inhibitory effect on pathogen. In fact, alginooligosaccharide derived from algal cell walls was found to be potential signals for the induction of chitinase expression in plant cells [Fry et al., 1993b]. After being stimulated with alginooligosaccharide, the phytoalexin accumulation in soybean cotyledon was observed, and

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such elicitor activity was related to the DP of alginooligosaccharides. The maximal production of phytoalexin was obtained when the alginooligosaccharide with DP of 6 was applied on the wound cotyledon (unpublicated data). The suppression of PAL gene expression has been proved to lead the increased fungal disease susceptibility in plants [Maher et al., 1994]. Thus the increase in PAL activity was a direct response of host plants to attempted penetration by the pathogens, which was associated with the resistance of plant to fungal diseases [Hahlbrock and Scheel, 1989; Shiraishi et al., 1995]. When alginooligosaccharide was applied on the surface of the wound soybean cotyledon, the increase in PAL could be detected instantly, and the peak value of PAL activity occurred at 1.5-2 h. After that there was another peak PAL activity appeared at 4-5 h.

6. CONCLUSION Polysaccharides can be degraded by different enzymes sourced from various living organisms, including microorganism, plant and animal. The carbohydrate-degrading enzymes can be classified as hydrolyase and lyase, which catalyses the degradation of polysaccharides in endo- or exo-type respectively. The enzymes degrading polysaccharides showed the specificity on their substrates. Some oligosaccharides, produced by enzymatic degradation of polysaccharides sourced from main constituents of pathogen or plant cell walls and virulence factors of exopolysaccharides of pathogen, have been shown biological activity, such as elicitor activity and antimicrobial activity. Such bioactivity on plant cells greatly depends on the molecular weight or degree of polymerization, the charge distribution, the arrangement of branch side chain, and terminal groups, etc.. The biodegradation of polysaccharides and their products, bioactive oligosaccharides, play a significant role in protect phytopathogen infection.

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In: Environmental Biodegradation Research Focus Editor: B. Y. Wang, pp. 49-76

ISBN: 978-1-60021-904-7 © 2007 Nova Science Publishers, Inc.

Chapter 2

BIODEGRADATION OR METABOLISM OF BISPHENOL A IN THE ENVIRONMENT Jeong-Hun Kang∗ and Yoshiki Katayama Department of Applied Chemistry, Faculty of Engineering, Kyushu University, Nishi-Ku, Fukuoka City, Japan

ABSTRACT Recently, there has been increasing interest in the effects of endocrine disruptors on organisms. Bisphenol A (BPA; 2,2-bis(4-hydroxyphenyl)propane; CAS Registry No. 8005-7) is an endocrine disruptor with estrogenic activity and acute toxicity to aquatic organisms. BPA is made by combining acetone and phenol and is used mainly as a material for the production of epoxy resins and polycarbonate plastics. Due to intensified usage of these products, exposure of organisms to BPA via several routes, such as the environment and the food chain, has increased. BPA contamination in the environment occurs through several routes, such as migration from human wastes and effluent from wastewater treatment plants. BPA exposed to the environment can be biodegraded or metabolized by microorganisms (bacteria, fungi and plankton), plants, invertebrates and vertebrates (fish, amphibians and mammals). Biodegradation or metabolism is a very important step for removing or detoxifying BPA in the environment or organisms. Although some metabolites of BPA may exhibit enhanced estrogenicity or toxicity, in general, BPA biodegradation or metabolism by organisms leads to detoxication of BPA. However, excessive BPA doses cause bioaccumulation if detoxification pathways are saturated. In this chapter we describe 1) contamination routes of BPA, 2) biodegradation or metabolism of BPA by organisms, and 3) bioaccumulation of BPA in organisms, with the main subject of this chapter being the biodegradation or metabolism of BPA by organisms.

∗

Corresponding author: Jeong-Hun Kang; Telephone number: 81-92-802-2849; Fax: 81-92-802-2849; E–mail address: [email protected]

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Key words: bisphenol A, biodegradation, metabolism, environment, bioaccumulation.

1. INTRODUCTION Bisphenol A (BPA; 2,2-bis(4-hydroxyphenyl)propane；CAS Registry No. 80-05-7) is a known endocrine disruptors (Krishnan et al., 1993) that is acutely toxic to aquatic organisms in the 1000–10,000 μg/l range for both freshwater and marine species (Alexander et al., 1988). BPA is an organic compound composed of two phenol rings connected by a methyl bridge, with two methyl functional groups attached to the bridge (Figure 1). BPA is used as a material for the production of phenol resins, polyacrylates and polyesters, but mainly for the production of epoxy resins and polycarbonate plastics. The epoxy resins are used as foodcontact surface lacquer coatings for cans, metal jar lids, protective coatings and finishes, automobile parts, adhesives, aerospace applications, and as a coating for PVC pipes. The polycarbonate plastics are used in compact disk manufacturing, automotive lenses, household appliances, food packaging, and plastic bottles (Staples et al., 1998). Due to increased use of products based on epoxy resins and polycarbonates, exposure of organisms to BPA via several routes, such as the environment and the food chain, has increased (Kang et al., 2006b). Thus, it is difficult for organisms (microorganisms, plants, invertebrates, and vertebrates) to escape from the toxic and endocrine disrupting effects of BPA. The transport potential of BPA to air is much lower (<0.0001%; 2.48e-4 to 0.351 ng/m3) than that to water (about 30%) or soil (about 68%) (Staples et al., 1998). The soil adsorption coefficient (Koc) values of BPA ranged from 314 to 1524 when calculated using a water solubility of 120 mg/l and an octanol-water partition coefficient (Kow) of 3.32 (Howard, 1989). These absorption values mean that BPA released into ground or surface waters can be absorbed by soils or sediments. BPA contamination in the environment occurs through several routes, such as by migration from human wastes and effluent from wastewater treatment plants. Photodegradation is one route of BPA removal from the environment. However, a major method of BPA elimination from the environment and from organisms is biodegradation or metabolism of BPA. In this chapter we describe 1) contamination routes of BPA, 2) biodegradation or metabolism of BPA by organisms, and 3) bioaccumulation of BPA in organisms. However, the biodegradation or metabolism of BPA by organisms is the main subject of this chapter.

2. CONTAMINATION ROUTES OF BPA IN THE ENVIRONMENT BPA discharge into the aquatic environment occurs not only from the migration of BPAbased products into rivers and marine waters but also through effluent from wastewater treatment plants and landfill sites. The primary route of BPA contamination in the aquatic environment is effluent from several plants and landfill sites. As regards soil, although BPA contamination occurs through migration from human wastes, the absorption of BPA from the aquatic environment to soil is a major route of BPA contamination (Kawahata et al., 2004).

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2.1. Effluent from wastewater treatment plants and landfill sites Biotic and abiotic methods are used for treating wastewater containing BPA. BPA is found in wastewater from paper and plastic production plants and domestic sewage treatment plants, as it is not completely removed during treatment (Fürhacker et al., 2000, Lee and Peart, 2002b; Rigol et al., 2002; Fuerhacker, 2003; Quinn et al., 2003). There is no addition of BPA during the manufacturing of paper. However, phenolic resins are used as binding agents in printing inks and polyester for coating of paper. As mentioned in the Introduction, BPA is used in the production of phenolic resins and polyester. Several studies have reported that wastewaters from kraft pulp, printing paper, and packing-board paper plants contain high concentrations of BPA (Quinn et al., 2003; Rigol et al., 2002). Moreover, BPA was found in wastewater from wastepaper recycling plants, which use thermal paper and/or printing paper as a raw material (Rigol et al., 2002; Fukazawa et al., 2001, 2002). High levels of BPA were identified in leachates from waste landfills (Yamada et al., 1999; Behnisch et al., 2001; Lee and Peart, 2000a; Yamamoto et al., 2001; Filho et al., 2003; Urase and Miyashita, 2003; Asakura et al., 2004). Yamamoto et al. (2001) reported levels of BPA in leachates from a hazardous waste landfill ranging from 1.3 to 17,200 μg/l (average 269 μg/l). Of course, the concentrations of BPA in effluent are considerably lower because these leachates are discharged following treatment. For example, Yamada et al. (1999) found that the levels of BPA in four landfill leachates ranged from 15 to 5400 μg/l, but ranged from 0.5 to 5.1 μg/l in the effluents after treatment. However, effluents that contain BPA after leachate treatment are known to be a source of BPA contamination in the environment (Yamada et al., 1999; Behnisch et al., 2001; Yamamoto et al., 2001; Filho et al., 2003; Asakura et al., 2004; Zha and Wang, 2006).

2.2. BPA migration from BPA-based products BPA can be leached to water from plastic wastes. A Japanese study reported that the levels of BPA leached from waster plastics, such as polyvinyl chloride (PVC) products and synthetic leather, ranged 1980–139,000 μg/kg (Yamamoto and Yasuhara, 1999). Fromme et al. (2002) reported BPA levels of between 610 and 1110 μg/kg dry weight in liquid manure samples. A possible source of the high BPA levels in liquid manure is migration from the inner surface-coating of tanks. Moreover, BPA migration from PVC hoses used for drainage, watering and sprinkling has been reported to range from 4–1730 μg/l (Yamamoto and Yashuhara, 2000). BPA contamination in soil can be positively correlated with human densities because of an increase in BPA pollution by human wastes such as domestic and/or industrial wastes (Kawahata et al., 2004).

2.3. BPA levels in river water and sediment According to studies conducted in the USA, Germany, Japan, Spain, China, Italy, and the Netherlands (Table 1), BPA levels in river water were 8 μg/l or less, except in one river water sample (21 μg/l) (Belfroid et al., 2002). Although BPA levels in river water near wastewater treatment plants or landfills can be high, downstream they may decrease to zero or to very

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low levels (Kim et al., 2004; Kang and Kondo, 2006a). This may be a result of biodegradation by microorganisms or a dilution effect. BPA levels in sediment ranged from <0.5 to 1630 μg/kg (Table 1). Generally, BPA levels in sediments are higher than those in surface river water. BPA in surface river water can be absorbed to sediments, based on the Koc values (314–1524) for BPA (Howard, 1989). Absorption of BPA into sediments increases with a decrease in temperature in the range of 4– 30℃ and with a decrease in pH in the range of 2.6–7 (Zeng et al., 2006). Moreover, BPA in anaerobic or semi-aerobic sediment environments can persist for a prolonged periods of time (Voordeckers et al., 2002), leading to higher BPA levels in sediments than in surface waters. Table 1. BPA levels in water and sediment samples BPA levels Sediment Countries References River water (μg/l) (μg/kg/dry weight) <0.05-1.51 Spain Céspedes et al., 2006 0.01-1.4 Japan JMC, 1999 0.02-0.15 Japan Takahashi et al., 2003 <0.2-1.9 Japan Matsumoto, 1982 <0.09 Japan Matsumoto et al., 1977 <0.5-0.9 Japan Kang and Kondo, 2006a <0.012-0.33 (21 in The Netherlands Belfroid et al., 2002 one sample) <1-8 USA Staples et al., 2000 0.03-0.083 China Jin et al., 2004 0.0005-0.014 Germany Kuch and Ballschmiter, 2001 0.21 Italy Urbatzka et al., 2007 <0.0088-1 <1.1-43 The Netherlands Vethaak et al., 2005 <0.005-0.08 <0.5-13 Japan Kawahata et al., 2004 0.02-0.03 0.11-48 Japan Hashimoto et al., 2005 <0.05-0.272 <0.5-15 Germany Bolz et al., 2001 0.004-0.092 10-380 Germany Stachel et al., 2003 0.009-0.776 66-343 Germany Heemken et al., 2001 0.0005-0.41 10-190 Germany Fromme et al., 2002 5-1,630 Germany Stachel et al., 2005 0.6-3.8 China Peng et al., 2006a 97-126 China Song et al., 2006

2.4. Chlorinated derivatives of BPA Chlorinated derivatives of BPA are by-products of the reaction between BPA and free chlorine from sodium hypochlorite, which is widely used as a bleaching agent in paper-pulp factories and a disinfectant in wastewater treatment (Hu et al., 2002; Yamamoto and

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Yasuhara, 2002; Gallard et al., 2004). As BPA contains two phenol rings connected by a methyl bridge, its phenolic structure reacts strongly with halogens, such as chlorine and bromide. The reactivity of BPA with chlorine is greater than that of phenol (Gallard et al., 2004). Since sufficient chlorine levels are needed to produce the chlorinated derivatives of BPA, their production in freshwater or marine waters is limited. However, free chlorine from sodium hypochlorite used in paper-manufacturing plants is present at adequate concentrations for the production of chlorinated derivatives of BPA. The chlorinated derivatives of BPA are 3-CIBPA [3-chlorobisphenol A; 2-(3-chloro-4hydroxyphenyl)-2-(4-hydroxyphenyl)propane], 3,5-diClBPA [3,5-dichlorobisphenol A; 2(3,5-dichloro-4-hydroxyphenyl)-2-(4-hydroxyphenyl)propane], 3,3´-diClBPA [3,3´-dichlorobisphenol A; 2,2-bis(3-chloro-4-hydroxyphenyl)propane], 3,3´,5-triClBPA [3,3´,5-trichlorobisphenol A; 2-(3,5-dichloro-4-hydroxyphenyl)-2-(3-chloro-4-hydroxyphenyl)propane], and 3,3´,5,5´-tetraClBPA [3,3´,5,5´-tetrachlorobisphenol A; 2,2-bis(3,5-dichloro-4-hydroxyphenyl)propane] (Figure 1). Moreover, polychlorinated phenoxyphenols have been identified in chlorinated BPA solutions (Hu et al., 2002). All chlorinated derivatives showed a more potent estrogenic activity (3–38-fold) than BPA alone in an agonist assay using a two-hybrid yeast system: the estrogenicity of 3,3´diClBPA, in particular, was 38-fold higher than that of BPA (Fukazawa et al., 2002). Among the chlorinated derivatives of BPA, the affinities of 3-ClBPA and 3,3´-diClBPA for the human α-estrogen receptor (ER) were higher than that of BPA (Takemura et al., 2005). Regarding the biodegradation of BPA and its chlorinated derivatives; BPA and 3-ClBPA are rapidly biodegraded (>50% after 1 day), but 3,3’,5-triClBPA and 3,3´,5,5´-tetraClBPA were scarcely biodegraded after 7 days under the same conditions, using the supernatant of the activated sludge in waste-paper recycling plants. Moreover, biodegradation of 3,3’diClBPA was about 50% after 7 days (Fukazawa et al., 2001). Their resistance to biodegradation indicates that these chlorinated derivatives may be stable in the aquatic environment.

2.5. BPA production from TBBPA Tetrabromobisphenol A [4,4´-isopropylidenebis(2,6-dibromophenol)] (TBBPA) is the most widely used flame-retardant in the production of many plastic polymers and electronic circuit-boards (Figure 2). Its toxic effects and metabolism in humans and wildlife have been reviewed previously (Darnerud, 2003; Hakk and Letcher, 2003). Anaerobic microorganisms in the sediments can dehalogenate TBBPA to BPA, but very little biodegradation of TBBPA was found under aerobic conditions (Ronen and Abeliovich, 2000; Arbeli and Ronen, 2003; Hakk and Letcher, 2003; Brenner et al., 2006). On the other hand, BPA is degraded slowly (e.g. over 3 months) under anaerobic conditions (Ronen and Abeliovich, 2000; Kang and Kondo, 2002a). Therefore, dehalogenation of TBBPA to BPA under anaerobic conditions may be an important source of BPA contamination in sediments (Voordeckers et al., 2002; Arbeli and Ronen, 2003; Hakk and Letcher, 2003).

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Jeong-Hun Kang and Yoshiki Katayama

Figure1. Chemical structure of BPA and its chlorinated derivatives.

Figure 2. Chemical structure of TBBPA, tri-BBPA, di-BBPA and mono-BPA.

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3. BIODEGRADATION AND METABOLISM OF BPA IN THE ENVIRONMENT BPA can be degraded or metabolized by microorganisms (bacteria, fungi, and planktons) (Table 2), plants, invertebrates, and vertebrates (fish, amphibians, and mammals). Biodegradation or metabolism is a very important step in removing or detoxifying BPA in the environment or in organisms (Kang et al., 2006a).

3.1. Bacteria 3.1.1. Bacteria capable of biodegrading BPA Many bacteria capable of BPA biodegradation have been identified from soils (Sasaki et al., 2005; Masuda et al., 2007), river waters (Ike et al., 2000; Kang and Kondo, 2002a, b; Kang et al., 2004), seawater (Sakai et al., 2007), and wastewater treatment plants (Lobos et al., 1992; Spivack et al., 1994) (Table 2). Wastewater treatment plants use bacteria to remove BPA from wastewater. Moreover, bacteria capable of biodegrading BPA are distributed in river waters and the half-lives for BPA biodegradation average below 5 days (Dorn et al., 1987; Jin et al., 1996; Ike et al., 2000; Klecka et al., 2001; West et al., 2001; Kang and Kondo, 2002a, b; Kang et al., 2004; Suzuki et al., 2004; Kang and Kondo, 2005). Although there are many bacteria capable of degrading BPA in river waters, bacteria with high BPA biodegradability are limited (Jin et al., 1996; Kang and Kondo, 2002a). For example, Kang and Kondo (2002a) found that most bacteria (10 out of 11) isolated from three river waters had BPA biodegradability, but there were differences in the removal rates of BPA (18 to 91%) and only two strains (a Pseudomonas sp. and a Pseudomonas putida strain) showed high BPA biodegradability (about 90%). The primary factor responsible for BPA degradation in soil is microorganisms, mainly bacteria (Fent et al., 2003). Interestingly, a strain (Pseudomonas monteilii) isolated from field soil had good tolerance to and biodegradation ability for higher levels of BPA (500,000–1,000,000 μg/kg) than those showing an acute toxicity to aquatic organisms (1000–10,000 μg/l) (Masuda et al., 2007). Bacteria with high BPA biodegradability may be useful for the fast purification of the environments contaminated by BPA. Routes of BPA metabolism by bacteria were determined by using a Gram-negative bacterial strain MV1 isolated from the enriched sludge of a BPA wastewater treatment plant. The MV1 strain utilized BPA as the sole carbon and energy source, and major and minor pathways of BPA metabolism were identified. The major pathway produced two primary metabolites, 4-hydroxyacetophenone and 4-hydroxybenzoic acid, and the minor pathway also produced two primary metabolites, 2,2-bis(4-hydroxyphenyl)-1-propanol and 2,3-bis(4hydroxyphenyl)-1,2-propanediol: the 4-hydroxybenzoic acid was formed from 4hydroxyacetophenone by oxidative rearrangement. Total carbon analysis for BPA showed that 60% of the carbon was mineralized to CO2, 20% was associated with the bacterial cells and 20% was converted to soluble organic compounds (Lobos et al., 1992; Spivack et al., 1994). A similar metabolic pathway for BPA was found in several other studies (Suzuki et al., 2004; Sasaki et al., 2005a; Masuda et al., 2007; Sakai et al., 2007; Zhang et al., 2007).

56

Jeong-Hun Kang and Yoshiki Katayama Table 2. Microorganisms capable of degrading bisphenol A

Microorganisms Bacteria

Fungi

Strains Sphingomonas paucimobilis FJ-4 Pseudomonas sp.

Sources Sludge in sewage plant River water

Pseudomonas putida

River water

Streptomyces sp Sphingomonas bisphenolicum strain AO1 Archromobacter xylosoxidans strain B-16 Sphingomonas sp. strain BP-7 Pseudomonas monteilii N502 Pleurotus ostreatus O-48 Phanerochaete chrysosporium ME-446 Trametes versicolor IFO7043 Trametes villosa

River water Soil of Vegetable farm Compost leachate of municipal solid waste Seawater

Sakai et al., 2007

Field soil

Masuda et al., 2007

NR a NR

Hirano et al., 2000 Tsutsumi et al., 2001; Suzuki et al., 2003 Tsutsumi et al., 2001; Suzuki et al., 2003 Fukuda et al., 2001; Uchida et al., 2001 Yim et al., 2003

Aspergillus fumigatus Fusarium sporotrichioides NFRI-1012 Fusarium moniliforme 2-2 Aspergillus terreus MT-13 Emericella nidulans MT98 Stereum hirsutum Heterobasidium insulare Coriolopsis polyzona

Planktons

Family Chaetomiaceae strain I-4 Chlorella fusca var. vacuolata Nannochloropsis sp. Chaetocero gracilis

a

NR, no reference.

NR NR

References Ike et al., 1995, 2002 Kang and Kondo, 2002a Kang and Kondo, 2002a Kang et al., 2004 Sasaki et al., 2005a; Oshiman et al., 2007 Zhang et al., 2007

Provision from institute Provision from institute Provision from institute Provision from institute Provision from institute Provision from institute Provision from institute Provision from institute Soil

Saito et al., 2004

NR

Hirooka et al., 2003

Provision from institute Provision from institute

Ishihara and Nakajima, 2003 Ishihara and Nakajima, 2003

Chai et al., 2005 Chai et al., 2005 Chai et al., 2005 Chai et al., 2005 Lee et al., 2005 Lee et al., 2005 Cabana et al., 2007

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The cytochrome P450 monooxygenase system is involved in BPA biodegradation (Table 3). Addition of cofactors (e.g., NADH, NAD+, NADPH, or NADP+) relating to the oxidation by cytochrome P450s leads to an increase in the BPA-degradation activity, but addition of metyrapone, a specific inhibitor of cytochrome P450, inhibits BPA biodegradation in Sphingomonas sp. strain AO1 cells (Sasaki et al., 2005a). NADH is a more effective electron donor for cytochrome P450s than NADPH. Cytochrome P450 accelerates BPA degradation in the presence of ferredoxin, ferredoxin reductase, and NADH, but there is no degradation in their absence (Sasaki et al., 2005b). Moreover, microbial peroxidase (Coprinus cinereus peroxidase) can degrade BPA. The optimal pH and temperature are 10 and 40℃, respectively. The BPA removal by microbial peroxidase depends on the H2O2 concentration (Sakurai et al., 2001). BPA biodegradation by bacteria eliminates the toxic or estrogenic effects of BPA. Ike et al. (2002) reported that, among 4 metabolites, only 4-hydroxyacetophenone showed slight estrogenic activity.

3.1.2. Factors influencing BPA biodegradation by bacteria Temperature and pH. Temperature is a very important factor both in cell growth and in BPA biodegradation (Kang and Kondo, 2002b; Zhang et al., 2007). For example, we found that the half-lives for BPA biodegradation in 15 river water samples averaged 4 days and 7 days at 30℃ and 20℃, respectively, but only 20% of spiked BPA was biodegraded at 4℃ over 20 days (Kang and Kondo, 2002b). Bacteria can degrade BPA over a wide pH range (e.g., 5–9), but the best optimal pH for BPA biodegradation was 7 to 8 (Masuda et al., 2007; Zhang et al., 2007). Bacterial counts. Our previous study suggested that bacterial counts have an influence on BPA biodegradation (Kang and Kondo, 2002b). In contrast, Klecka et al. (2001) reported that BPA biodegradation was not correlated to bacterial counts. These differences may be due to the number of 1) bacteria capable of performing complete BPA biodegradation or mineralization and 2) bacteria previously exposed to BPA. Bacteria previously exposed to BPA may have developed BPA biodegradation ability. Moreover, bacteria with high BPA biodegradation and mineralization abilities were mainly isolated from river waters with high bacterial counts (Jin et al., 1996). Aerobic and anaerobic conditions. There is a significant difference between BPA biodegradation between under aerobic conditions and under anaerobic conditions. BPA in river waters is biodegraded under aerobic conditions but not under anaerobic conditions. Our group found that BPA in the spiked river samples was rapidly biodegraded under aerobic conditions (>90%), but hardly any decrease in BPA was found under anaerobic conditions (<10 %) over 10 days (Kang and Kondo, 2002a). Similar results were obtained for seawater and marine sediment (Ying and Kookana, 2003). In addition, BPA in an anaerobic slurry was not biodegraded even after 3 months of incubation (Ronen and Abeliovich, 2000). In soil, BPA is biodegraded under aerobic conditions and the final product of BPA in is CO2 (Fent et al., 2003), but no BPA biodegradation occurred under anaerobic conditions, even after 70 days of incubation (Ying and Kookana, 2005). These results may mean that anaerobic bacteria have no or little BPA biodegradability.

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Nutrients. Bacteria can utilize BPA as their sole carbon and energy source, but the addition of nutrients (e.g., peptone and glucose) stimulates BPA-biodegrading activity by increasing bacterial growth (Sasaki et al., 2005a; Sakai et al., 2007) and increasing the resistance of bacteria to the toxic effects of BPA (Sasaki et al., 2005a; Zhang et al., 2007). The effects of nutrients on BPA biodegradation differ. For example, Sphingomonas sp. strain BP-7 can biodegrade BPA in the presence of peptone, but the presence of glucose has no effect (Sakai, et al., 2007). On the other hand, the initial BPA biodegradation rate in P. monteilii strain N-502 was accelerated by the addition of Ca2+, Mg2+, and folic acid (Masuda et al., 2007). BPA has a longer residence time before biodegradation occurs in seawater than in river water (Ying and Kookana, 2003; Kang and Kondo, 2005). Seawater is a nutrient-poor environment, and the environments can have an influence on both BPA biodegradation and bacterial growth (Sakai et al., 2007). Moreover, BPA degradation in seawater may be influenced more by chemical degradation than biological degradation, while organisms such as bacteria and flagellates can have an important effect on the chemical degradation of BPA (Kang and Kondo, 2005). BPA intermediates. 4-Hydroxyacetophenone is a biodegradation intermediate of BPA. Its accumulation inhibits bacterial growth and BPA biodegradation (Sasaki et al., 2005a; Sakai et al., 2007). However, biodegradation of both BPA and 4-hydroxyacetophenone is accelerated by the addition of nutrients (Sasaki et al., 2005a).

3.2. Fungi 3.2.1. Fungi capable of biodegrading BPA Many fungi can degrade BPA (Table 2), but fungi with high BPA biodegradability are also limited (Chai et al., 2005; Yim et al., 2003). For example, Chai et al. (2005) found that among 26 fungi strains, 11 strains could biodegrade BPA at >50% BPA and 4 strains (Fusarium sporotrichioides NFRI-1012, Fusarium moniliforme 2-2, Aspergillus terreus MT13 and Emericella nidulans MT-98) were more effective at BPA biodegradation.

3.2.2. BPA degradation by lignin-degrading enzymes, manganese peroxidase and laccase BPA degradation by fungi is caused mainly by lignin-degrading enzymes such as manganese peroxidase (MnP) and laccase, which are produced by white rot basidiomycetes fungi (Table 3) (Hirano et al., 2000; Fukuda et al., 2001; Tsutsumi et al., 2001; Uchida et al., 2001; Suzuki et al., 2003; Kim and Nicell, 2006; Cabana et al., 2007). MnP is a heme peroxidase that oxidizes phenolic compounds in the presence of Mn2+ and H2O2. Laccase is a multicopper oxidase that catalyzes one-electron oxidation of phenolic compounds by reducing oxygen to water (Reinhammar, 1984). In the case of laccase, BPA biodegradation is faster in the presence of mediators such as 1-hydroxybenzotriazole (HBT) and 2,2’-azino-bis(3ethylbenzthiazoline-6-sulfonate) than in laccase alone: 2,2’-azino-bis(3-ethylbenzthiazoline6-sulfonate) is more effective than HBT (Tsutsumi et al., 2001; Cabana et al., 2007). The optimum pH and temperature for BPA degradation by laccase are 5–6 and 40–50℃,

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respectively (Tanaka et al., 2001; Kim and Nicell, 2006; Cabana et al., 2007). Interestingly, the laccase from Trametes sp. can degrade BPA, but no BPA degradation was observed by the laccase prepared from Pycnoporus coccineus (Tanaka et al., 2001). This means that the activity of laccase for BPA differs between the fungal species. Table 3. Bisphenol A-degrading enzymes in organisms Enzymes Manganese (MnP)

peroxidase

Laccase

Peroxidase

Tyrosinase Cytochrome P450

UDPglucuronosyltransferase (UGT)

Sources Fungi (Pleurotus ostreatus O-48, Phanerochaete chrysosporium ME-446, Trametes versicolor IFO-7043, Phanerochaete chrysosporum ME-446, and Trametes versicolor IFO-6482) Fungi (Phanerochaete chrysosporium ME-446, Trametes versicolor IFO-7043, Trametes villosa, Phanerochaete chrysosporum ME446, and Trametes versicolor IFO-6482) Bacteria (Coprinus cinereus) Plant [soybean and horseradish (Armoracia rusticana)] Plant (mushroom) Bacteria (Sphingomonas bisphenolicum strain AO1) Mammals (mouse and rat)

References Hirano et al., Tsutsumi et al., Suzuki et al., 2003

2000; 2001;

Tsutsumi et al., 2001; Fukuda et al., 2001; Uchida et al., 2001; Suzuki et al., 2003; Kim and Nicell, 2006 Sakurai et al., 2001; Caza et al., 1999; Sakuyama et al., 2003 Yoshida et al., 2002 Sasaki et al., 2005a

Fish [carp (Cyprinus carpino)]

Atkinson and Roy, 1995a; 1995b; Yoshihara et al., 2001 Yokota et al., 2002

Mammals (mouse, rat, and human)

Yokota Cappiello

et et

al., al.,

1999; 2000;

Matsumoto et al., 2002; Strassburg et al., 2002 Sulfotransferase

Mammal (human)

Suiko et al., 2000; Nishiyama et al., 2002

BPA degradation by MnP and laccase removes its estrogenic activity (Hirano et al., 2000; Fukuda et al., 2001; Tsutsumi et al., 2001; Uchida et al., 2001; Suzuki et al., 2003; Saito et al., 2004; Lee et al., 2005; Cabana et al., 2007). BPA degraded by MnP is metabolized to phenol, 4-isopropenylphenol, 4-isopropylphenol and hexestrol (Hirano et al., 2000). On the other hand, polymerization of BPA to form oligomers is included in the BPA metabolism step using laccase, followed by either the addition of phenol moieties or the degradation of the oligomers to release 4-isopropenylphenol. Moreover, a BPA dimmer, 5,5’-bis-[1-(4-hydroxyphenyl)-1-methyl-ethyl]-bisphenyl-2,2’-diol, from oligomers as a high molecular weight compound, was identified from BPA biodegradation by laccase (Uchida et al., 2001). These oligomers are produced by the formation of C–C or C–O bonds (Cabana et al., 2007).

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3.3. Planktons Algae have the potential possibility to remove various pollutants including BPA (Nagase et al., 1997). Algae use CO2 as a carbon source and grow photoautotrophically. These characteristics mean low removal rates of BPA under dark conditions. For example, the removal rate of BPA by a green alga Chlorella fusca var. vacuolata was >85% under light conditions after 120 h (>85%), but was <30% under dark conditions when 40 μM of BPA was added into the medium (Hirooka et al., 2003). Moreover, algae participates in the production of reactive oxygen species (ROS) [e.g., hydroxyl radicals (OH·)] under light conditions, and product formation increases in the presence of humic acid and Fe3+ (Peng et al., 2006b): ROS enhances BPA degradation (see 6. Photodegradation of BPA in the environment). Monohydroxybisphenol A was identified as an intermediate of BPA biodegradation by C. fusca (Hirooka et al., 2003; 2005). BPA biodegradation by algae leads to the removal of estrogenic activity of BPA (Hirooka et al., 2005).

3.4. Plants BPA metabolism by plants has been examined by using plant cell suspension cultures (Nakajima et al., 2002, 2004; Schmidt and Schuphan, 2002; Chai et al., 2003), plant enzymes (Sakurai et al., 2001; Xuan et al., 2002; Yoshida et al., 2002; Kang et al., 2006c), or direct absorption of BPA into plants (Nakajima et al., 2002; Noureddin et al., 2004; Kang and Kondo, 2006b). Plants can rapidly absorb BPA through their roots and metabolize it to several glycosidic compounds: the glycosylation of BPA occurs mainly in the roots (Nakajima et al., 2002; Noureddin et al., 2004; Kang and Kondo, 2006b). Nakajima et al. (2002) found that the BPA absorbed through root systems was metabolized to its β-glucoside and the metabolites were translocated to the leaves. On the other hand, Noureddin et al. (2004) suggested that BPA metabolites are detected at ca. 10% in the roots, some in the stems, but none in the leaves. These results show that the distribution of BPA and its metabolites in plants may differ between plant species. Glycosylation of BPA is regarded as the main route of BPA metabolism in plants. The metabolites from glycosylation of BPA have been well studied by Nakajima and his colleagues (Nakajima et al., 2004). They identified two major products, BPA mono-O-β-Dgentiobioside and the trisaccharide BPA mono-O-β-D-glucopyranosyl-(1→4)-[β-Dglucopyranosyl-(1→6)] β-D-glucopyranoside, and two minor products, mono- and di- O-β-Dglucopyranosides. Moreover, no estrogenic activity was found in these compounds, meaning that the glycosylation of BPA by plants destroys the estrogenicity of the parent compound. Plant enzymes have an important effect on BPA degradation. BPA is degraded by crude enzymes prepared from vegetables and fruits (Xuan et al., 2002; Kang et al., 2006c). In particular, two oxidative enzymes, peroxidase and tyrosinase, are closely related to BPA degradation (Table 3) (Caza et al., 1999; Yoshida et al., 2001; 2002; Sakuyama et al., 2003). In plants, generally, physiological stress, wounding and microbial or viral infections result in increased activity of tyrosinase and peroxidase activity, for self-protection (Vámos-Vigyázó, 1981). In the case of BPA degradation by plant enzymes, the pH and temperature exert an

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important influence on the oxidative removal of BPA (Caza et al., 1999; Xuan et al., 2002; Yoshida et al., 2002; Sakuyama et al., 2003; Kang et al., 2006c). The optimal pH and temperature for the BPA removal by peroxidase are 7–8 and 20–25℃, respectively (Caza et al., 1999; Sakuyama et al., 2003). For BPA degradation by tyrosinase, the optimal pH and temperature are 7–8 and 40–45℃, respectively (Yoshida et al., 2002). The BPA oxygenation by peroxidase depends on the H2O2 and peroxidase concentrations (Caza et al., 1999; Sakuyama et al., 2003; Huang and Weber, 2005). Most metabolic products of BPA formed by peroxidase are polymers, and some 4-isopropenylphenol has also been identified (Sakuyama et al., 2003; Huang and Weber, 2005). The polymerization of BPA by peroxidase may occurs through C–O bonds (Huang and Weber, 2005). The main product of BPA obtained from the enzymatic oxygenation of tyrosinase was the monoquinone derivative of BPA. A small amount of the bisquinone derivative was also identified (Yoshida et al., 2002). Moreover, the oxidation products of BPA obtained by using the potato enzyme were 4[1-(4-hydroxyphenyl)-1-methyl-ethyl]-benzene-1,2-diol and 4[1-(4-hydroxyphenyl)-1methyl-ethyl]-benzene-1,3-diol. The BPA oxidized by plant enzymes lost its estrogenic activity (Schmidt and Schuphan, 2002; Xuan et al., 2002).

3.5. Invertebrates There is a paucity of data on BPA metabolism in invertebrates. A study of the toxicokinetics of BPA in the freshwater clam, Pisidium amnicum, showed that the half-lives for BPA at 1.8 and 11.6℃ were 221 and 43 h, respectively. The rapid half-life for BPA with an increased temperature may be related to an elevated metabolic rate (Heinonen et al., 2002).

3.6. Vertebrates 3.6.1. Fish In fish, two BPA metabolites (BPA sulfate and BPA glucuronide) were identified from zebrafish (Danio rerio) exposed to BPA (Lindholst et al., 2003). UDPglucuronosyltransferase (UGT) is an intrinsic membrane protein and its substrates include phenols, carboxylic acids, alcohols and amines, including xenobiotics such as 4hydroxybiphenyl and opioid compounds. Glucuronidation by UGT is an important process in the metabolism of xenobiotic and endogenous substances leading to the enhanced detoxification of these compounds (Tephly and Burchell, 1990; King et al., 1997). UGT can metabolize BPA to BPA glucuronide in fish. An increase in UGT activity for BPA was found in microsomes prepared from carp (Cyprinus carpino) intestine where the BPA glucuronide was excreted into the mucosal side of the proximal, middle and distal intestinal segments (Yokota et al., 2002; Daidoji et al., 2006). UGT activity in carp intestine and excretion of BPA glucuronide from carp intestine are lower in winter than in summer. The low UGT activity leads to the accumulation of BPA in intestinal tissues (Daidoji et al., 2006). Interestingly, in fish, BPA is metabolized to BPA glucuronide mainly in the intestine (Yokota et al., 2002), whereas the liver is the main metabolic organ in mammals (Yokota et al., 1999; Nishiyama et al., 2002).

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On the other hand, BPA sulfate is a minor metabolite. In zebrafish, BPA glucuronide levels in plasma and bile are 100 and 22,600 times higher, respectively, than those of BPA sulfate (Lindholst et al., 2003). Moreover, the metabolism levels for BPA may vary with fish species. For example, BPA metabolism is faster in zebrafish liver than in rainbow trout liver (Lindholst et al., 2001, 2003).

3.6.2. Amphibians A recent study reported that uptake rates of BPA and its tissue levels in frog tadpoles were higher at 19℃ than at 7℃ during a 96-h experiment, but this didn’t lead to an increase in bioconcentration factors because of its higher elimination rates (Honkanen and Kukkonen, 2006). 3.6.3. Birds From an administration study of 14C-BPA to quail embryos, strong labeling in the bile and the allantoic fluid was identified, meaning that BPA is metabolized and excreted by the embryos (Halldin et al., 2001). Generally, chicken embryos have a relatively high metabolic capacity during the first half of incubation, both in terms of cytochrome P450-catalysed reactions and conjugation (Dutton and Ko, 1966; Heinrich-Hirsch et al., 1990). In laying quail, 14C-BPA administered orally and intravenously was rapidly removed via bile and excreted in feces (Halldin et al., 2001). 3.6.4. Mammals 3.6.4.1. Metabolism of BPA in mammals BPA metabolism in mammary involves two pathways, glucuronidation and sulfation of BPA. BPA is glucuronided mainly by liver microsomes (Table 3). Glucuronidation is mediated by UGT2B1, an isoform of UGT, in the rat liver (Yokota et al., 1999). The UGT superfamily is classified into 2 families (UGT1 and UGT2) and 3 subfamilies (UGT1A, UGT2A, and UGT2B). Each subfamily also contains several isoenzymes. The UGT enzymes are expressed not only in the liver but also in extrahepatic tissues (Kiang et al., 2005). Metabolism of BPA is faster in female rats than in male rats and the relative expression level of UGT2B1 mRNA is also higher in female rats than in male rats (Takeuchi et al., 2004a). Gender differences in serum BPA concentrations of adult humans may be caused by differences in the androgen-related metabolism of BPA (Takeuchi et al., 2002; 2004b). Higher serum BPA concentrations and lower BPA glucuronidation rates in ovariectomized rats were identified in the presence of than in the absence of testosterone propionate (TP) than in its absence. UGT2B1 mRNA was also lower in the TP injection group than in the non-TP injection group, meaning a negative effect of androgen on UGT2B1 (Takeuchi et al., 2006). Hepatic glucuronidation is slightly lower in pregnancy than in nonpregnancy because UGT levels decrease during pregnancy (Inoue et al., 2004). Moreover, the UGT levels in the human fetal liver are lower than those in the adult liver (Cappiello et al., 2000; Matsumoto et al., 2002; Strassburg et al., 2002). Activity of UGT toward BPA and its protein and mRNA contents were not detected in the fetal rat liver (Matsumoto et al., 2002). In addition, a continuous exposure to BPA leads to a decrease in the expression levels of UGT2B1 in male

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Wistar rats, but not in female rats (Shibata et al., 2002). On the other hand, human liver microsomes cannot glucuronidate BPA as extensively as the rat liver microsomes can (Elsby et al., 2001b). Microsomal cytochrome P450 enzymes in rat liver take part in the metabolism of BPA. An inhibitor of the cytochrome P450 system, SKF 525-A, inhibits the metabolism of BPA (Yoshihara et al., 2001). Cytochrome P450s can metabolize BPA into bisphenol-o-quinone via 5-hydroxy BPA and a bisphenol semiquinone (Atkinson and Roy, 1995a, b). On the other hand, BPA at concentrations of >20 mg/kg can act as an inhibitor for human hepatic cytochrome P450s activities (Hanioka et al., 1998; Niwa et al., 2000; Pfeiffer and Metzler, 2004). Among cytochrome P450s, BPA inhibits CYP2C8 and CYP2C19 in hepatic microsomes from humans (Niwa et al., 2000), and CYP2C11, CYP2C11/6, and CYP3A2/1 in hepatic microsomes from male Sprague-Dawley rats (Hanioka et al., 1998; Pfeiffer and Metzler, 2004). Sulfation of BPA by sulfotransferases in the liver is also included in the BPA metabolism pathway in mammals (Suiko et al., 2000; Nishiyama et al., 2002) (Table 3). Among human sulfotransferases, the simple phenol (P)-form phenol sulfotransferase (SULT1A1) (Suiko et al., 2000; Nishiyama et al., 2002) and thermostable phenol sulfotransferase (ST1A3) (Shimizu et al., 2002) show the sulfation of BPA. In addition, BPA glucuronide levels are higher in men than in women, but BPA sulfate levels are the opposite (Kim et al., 2003).

3.6.4.2. Excretion of BPA metabolites After glucuronidation or sulfation in the rat liver, the metabolites of BPA are excreted mainly into the bile (Inoue et al., 2001; 2004). However, higher levels of metabolites are eliminated to the veins in pregnant rats and in female rats. The venous excretion of metabolites increases 3-fold during pregnancy in comparison to during nonpregnancy (Inoue et al., 2004). There are differences by species or strain in the excretion of BPA metabolites. BPA glucuronide, a major metabolite of BPA, is mainly excreted via the bile into the feces in rats (Tominaga et al., 2006), but is excreted into the urine in primates (humans, monkeys, and chimpanzees) (Matsumoto et al., 2003; Yang et al., 2003; Calafat et al., 2005; Tominaga et al., 2006). Orally or subcutaneously administered BPA (10 mg/kg) in primates (monkeys and chimpanzees) is more easily absorbed than in rats and it takes a longer time to eliminate BPA in primates than in rats (Negishi et al., 2004; Tominaga et al., 2006). Faster elimination of BPA in the systemic circulation of rodents than that of primates may be caused by the hepatic blood flow-rate (Tominaga et al., 2006): the hepatic blood flow-rate is higher in rodents, mouse (90 ml/min/kg) and rat (55.2 ml/min/kg) than in primates, monkey (43.6 ml/min/kg), chimpanzee (25.5 ml/min/kg), and human (20.7 ml/min/kg) (Davies and Morris, 1993; Wong et al., 2004). For strain differences, F-344 rats excrete more BPA metabolites in urine than CD rats (Snyder et al., 2000). In Fisher 344 rats, female rats show higher concentrations of both BPA glucuronide and BPA sulfate than male rats (Pottenger et al., 2000). 3.6.4.3. Metabolites of BPA and their toxicity and estrogenicity BPA glucuronide is characterized as the major metabolite of BPA metabolism via the liver microsome pathway. Other metabolites such as BPA sulfate conjugate, BPA diglucuronide, 5-hydroxy BPA and the corresponding sulfate conjugate have also been reported (Atkinson and Roy, 1995a, b; Elsby et al., 2001a; Nakagawa et al., 2001; Shimizu et

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al., 2002). Jaeg et al. (2004) identified nine metabolites from the metabolism of BPA by CD1 mice liver microsomal and S9 fractions. The metabolites were isopropyl-hydroxyphenol, BPA glutathione conjugate, glutathionyl-phenol, glutathionyl 4-isopropylphenol and BPA dimmers (Jaeg et al., 2004). Some BPA metabolites show higher toxicity and estrogenicity than BPA. One BPA metabolite, bisphenol-o-quinone, binds DNA both in vitro and in vivo. These covalent modifications to DNA by in vivo exposure to BPA may be a factor in the induction of hepatotoxicity (Atkinson and Roy, 1995a, b). Moreover, the 4-methyl-2,4-bis(phydroxyphenyl)pent-1-ene (MBP) identified from BPA degradation by rat liver S9 fractions, microsomal and cytosolic fractions, showed 2- to 5-fold higher estrogenic activity than BPA (Yoshihara et al., 2001, 2004). In a study using medaka (Oryzias latipes) MBP showed higher toxicity during its early life stages and 250-fold higher estrogenic activity than BPA (Ishibashi et al., 2005). BPA glucuronide has lower estrogenicity than BPA. The estrogenicity of 5-hydroxy BPA is also less than that of BPA, but it is known as a weakly estrogenic compound (Nakagawa et al., 2001; Elsby et al., 2001a). The BPA sulfate shows no estrogenicity up to a concentration of 1 mM, but an increase in the levels of pS2 mRNA expression is found at a concentration of 1 μM of BPA (Shimizu et al., 2002).

5. BIOACCUMULATION Despite the low levels of BPA present in river water, BPA contamination has been found in aquatic organisms (Larsson et al., 1999; Belfroid et al., 2002; Takahashi et al., 2003; Kang and Kondo, 2006a). Takahashi et al. (2003) found that BPA levels were <0.2 μg/l in river water, but from 2–8.8 μg/kg wet weight in the periphytons and from 0.3–12 μg/kg wet weight in the benthos (periphytons are a complex matrix of algae and heterotrophic microbes that serve as an important food source for invertebrates and some fish; benthos are the organisms and habitat at the bottoms of lakes, rivers and creeks, or the sea floor). Moreover, levels of BPA in fish varied from 2 to 75 μg/kg dry weight (DW) in the liver and from 1 to 11 μg/kg DW in the muscle, but ranged from <0.01 to 0.33 μg/l in surface water (Belfroid et al., 2002). Bioaccumulation is closely related to BPA biodegradation. Excessive BPA doses lead to bioaccumulation if detoxification pathways are saturated (Upmeier et al., 2000). Bioconcentration factors for BPA in fish range from 5 to 68, suggesting a low potential for bioaccumulation (Staples et al., 1998; Lindholst et al. 2000). However, bioaccumulation can increase at lower temperatures because of lower elimination rate and metabolism of BPA. High bioconcentration factors (110–144) for BPA were found in the freshwater clam Pisidium amnicum at low temperatures (2–12℃) (Heinonen et al., 2002). Moreover, bioconcentration factors for salmon yolk-sac fry after 96 h of exposure ranged from 94 to 182, suggesting that the early life stages accumulate higher concentrations of BPA than the later stages (Honkanen et al., 2004). In spite of low uptake of BPA at a low temperature (7℃), the bioconcentration factors in frogs were 24 to 48% higher than those at a high temperature (19℃) (Honkanen and Kukkonen, 2006). Interestingly, bioconcentration factors in fish and frog cannot be influenced by exposure to BPA concentrations below the limits of toxicity (Honkanen et al., 2004; Honkanen and Kukkonen, 2006). However, solar ultraviolet-B radiation (280–320 nm)

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has no effect on the BPA accumulation in frogs (Koponen et al., 2007). In rats subcutaneously receiving 1 or 5 mg/day doses for 15 days, BPA accumulation into brown adipose tissue was found (Nunez et al., 2001). Bioaccumulation of BPA in organisms through the food chain is a very important factor. Ishihara and Nakajima (2003) found that the recovery of BPA from medium and from marine phytoplankton cells (Nannochloropsis sp.) was 11 and 46%, respectively, after spiking with 24 μM BPA. On the other hand, the recovery of BPA from medium and from zooplankton, such as Arterima sp. or Brachionus sp., was >80 and <7%, respectively. However, >40% of BPA was recovered from the zooplankton cells in the medium using a combination of the phyto- and zooplankton, meaning that BPA is accumulated in the zooplankton cells via the phytoplankton cells. Moreover, since BPA can persist longer in seawater than in freshwater, the possibility of BPA accumulation is higher in marine organisms than in freshwater organisms (Ying and Kookana, 2003; Kang and Kondo, 2005).

6. PHOTODEGRADATION OF BPA IN THE ENVIRONMENT Photodegradation is the main non-biological pathway of BPA breakdown in the environment and occurs by either photolysis or photooxidation. Photodegradation of BPA is slow in pure water, but can be accelerated in the presence of dissolved organic matter (DOM), including humic and fulvic acid (Chin et al., 2004; Peng et al., 2006b; Zhan et al., 2006), reactive oxygen species (ROS), including hydroxyl radicals (OH·), peroxyl radicals (ROO·) and singlet oxygen (1O2) (Sajiki and Yonekubo, 2002, 2003; Zhan et al., 2006), and/or ions, including ferric and nitrate ions (Zhou et al., 2004; Peng et al., 2006b; Zhan et al., 2006). DOM is widely present in surface water and can absorb irradiation and increase the transformation of BPA by generating reactive photooxidants, ROS and other non-ROS transients (Chin et al., 2004; Zhan et al., 2006). Iron is found in almost all aquatic environments and has a function in producing ROS by reaction with hydrogen peroxide (Johnson et al., 1997). Moreover, complexes of iron with DOM or ROS, such as Fe(III)-OH and Fe(III)-humic acid complexes, induce photodegradation of BPA (Zhou et al., 2004). The photodegradation products of BPA are phenol, 4-isopropylphenol and a semiquinone derivative of BPA (Howard, 1989). In air, the photo-oxidation half-life for BPA ranged from 0.74 to 7.4 h on the basis of hydroxyl radical reaction (Staples et al., 1998).

7. CONCLUSION Environmental BPA contamination occurs via several routes, such as migration from human wastes and effluent from wastewater treatment plants. The BPA exposed to the environment can be biodegraded or metabolized by microorganisms (bacteria, fungi, and planktons), plants, invertebrates, and vertebrates (fish, amphibians, and mammals). Biodegradation or metabolism is a very important step in removing or detoxifying BPA present in the environment or in organisms. Although some metabolites of BPA may enhance estrogenicity or toxicity, in general, BPA biodegradation or metabolism by organisms leads to

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detoxication of BPA. It should be noted that the fact that BPA can be biodegraded or metabolized by organisms does not mean that BPA has no estrogenic or toxic effect on organisms. Moreover, excessive BPA doses cause bioaccumulation if detoxification pathways are saturated. Bioaccumulation for several organisms increase at lower temperatures because of lower elimination rate and metabolism of BPA.

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In: Environmental Biodegradation Research Focus Editor: B. Y. Wang, pp. 77-100

ISBN: 978-1-60021-904-7 © 2007 Nova Science Publishers, Inc.

Chapter 3

FROM PLANTING TO HARVEST: ENVIRONMENTAL DISSIPATION OF THE HERBICIDE MOLINATE AND PROPOSAL OF A CLEAN-UP METHODOLOGY Célia M. Manaia1 and Olga C. Nunes2∗ 1

Escola Superior de Biotecnologia, Universidade Católica Portuguesa, 4200-072 Porto, Portugal. 2 LEPAE-Departamento de Engenharia Química, Faculdade de Engenharia, Universidade do Porto, 4200-465 Porto, Portugal

ABSTRACT The deliberated application and intensive use of pesticides in the environment has been leading to the contamination of air, soils, surface and ground water, and living organisms. The environmental contamination of the trophic chain with pesticides has serious negative impacts on the biological diversity and possible implications on the public health. The monitoring of environmental contamination with pesticides and the implementation of decontamination processes may contribute to minimize the impact of intensive agricultural practices. This study was conducted in a rice field situated in central Portugal, where molinate is supposed to contaminate surface and underground waters. Molinate content was monitored in water samples collected before, during and after molinate application and the results showed that molinate was dissipated in the environment, reaching concentrations of 3.9 µg l-1 in underground water and 15.8 µg l-1 in the river receiving tail waters. The feasibility of clean-up methodologies based on adsorption and/or biodegradation processes to remove molinate from these waters was assessed. At a laboratory scale, these clean-up processes led to reductions of the molinate concentration to values close to the legally recommended limits (< 2 µg l-1). Given the inability of the autochthonous microbiota to degrade molinate contaminating the agriculture effluents, the implementation of a biodegradation process requires the use of an exogenous molinate degrading culture. The aerobic biodegradation ∗ Send correspondence to Dr. Olga C. Nunes, Email: [email protected]

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Célia M. Manaia and Olga C. Nunes of molinate in the rice field waters was assayed using a defined mixed bacterial culture, previously isolated from an industrial effluent. This mixed culture is able to mineralize molinate under a wide range of operating conditions removing between 55 % (1 mM molinate at 15 °C) and 80 % (1 mM molinate and complex nutrients at 30 °C) of the initial total dissolved organic carbon. Given the low concentrations, and hence the low bioavailability, of molinate in agriculture effluents, the use of an adsorption step was considered a valuable auxiliary tool to improve the clean-up of contaminated waters. Resin Amberlite XAD-4 and activated carbon showed efficient molinate removal. The bio-regeneration of these materials, using the above mentioned mixed culture, permits the decontamination of the adsorbents, both for future re-use or for final disposal.

INTRODUCTION The use of pesticides to improve crops yields has become universal in last decades. Although it is arguable that pesticides have evident benefits in agriculture, these chemical substances may reach relatively high and dangerous concentrations in natural environments due to their deliberated application and intensive use. As for other xenobiotic compounds, the concentration of a pesticide in a specific site is not only determined by the amount and rate of application, but also by its miscibility, diffusiveness and rate of degradation under natural conditions. Thus, the physicochemical characteristics of both pesticide and receptor environment will determine its dissipation. Major modes of dissipation of pesticides include volatilization, run-off and leaching to surface and underground waters, adsorption onto soil particles, and (photo)chemical or biological degradation. Through these physicochemical processes the original compound can be diluted, concentrated or partially degraded. However, the mineralization of the parent pesticide, and/or of a thereof (photo)chemical transformation product is believed to be more rare in nature and may only occur if appropriate microbial populations, i.e., microorganisms able to (co-)metabolize the pesticide, are present. Moreover, even when adequate degraders are present, pesticide residues may accumulate in the environment due to its reduced bioavailability. When both complete (photo)chemical and biological degradation processes fail, the accumulation of recalcitrant transformation products, sometimes with increased toxicity, is unavoidable. The contamination of the trophic chain due to bioaccumulation and bioamplification processes may also occur with some hydrophobic pesticides and is another negative consequence of the use of these xenobiotics. These mechanisms are responsible, individually or in combination, for the pesticide contamination of air, surface and underground water, soil and biotic community reported worldwide, that occasionally exceeds the legal thresholds and that very often is believed to have negative impacts on the environmental health. The development of strategies to prevent environmental contamination (e.g. reduction of pesticide application, utilization of less toxic substances), and to clean-up contaminated sites is important. In contrast to what happens with industrial effluents, the implementation of agricultural wastewater treatment systems is difficult, due to the unconfined application of pesticides and the absence of effluent collectors. In this respect, the rice culture is an exception, because rice crops are confined in paddy fields irrigated with water. Given that the

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majority of the pesticides used in rice crops are applied to flooded fields, it is possible to promote the treatment of tail waters before its discharge into natural water streams.

THE HERBICIDE MOLINATE Molinate (S-ethyl-perhydroazepine-1-carbothioate) is one of the pesticides most extensively applied to rice fields worldwide. The physicochemical characteristics of this thiocarbamate herbicide are presented in Table 1. Molinate is a systemic herbicide applied to flooded fields, once or twice a year, during sowing, planting or when rice plants are small, to control the overgrowth of grass weeds, namely Echinochloa spp. and Glyceria spp. Molinate is the active compound of several commercial herbicides. The rate of molinate application usually varies between 2.5 and 5 kg ha-1, but rates up to 11 kg ha-1 are reported (Deuel et al., 1978; Ross and Sava, 1986), depending on the formulation, mode of application and agricultural practice. In a field study, in which molinate was applied twice in a total amount of about 7.5 kg ha-1, it was estimated that, on the day of the second application, 81 % of the molinate was in the water, 10 % in the soil, 9 % in the air and < 1 % in the vegetation (Ross and Sava, 1986). Due to its physicochemical properties, namely high water solubility and moderate volatility, molinate is described as being readily dissipated from paddy fields. Reported DT50 values of molinate vary between 2 and 5 days in water, 3 and 7 days in nonflooded soils and 5 and 35 days in flooded soils (SANCO, 2003). Molinate dissipation from rice fields occurs through volatilization, photodegradation, soil adsorption, microbial degradation, biota incorporation, leaching and surface run-off (Soderquist at al., 1977; Deuel et al., 1978; Imai and Kuwatsuka, 1982; Imai and Kuwatsuka, 1988; Ross and Sava, 1986; Carrasco et al., 1992; Johnson and Lavy, 1995; Albanis et al., 1998; Tsuda et al., 1998; Hernandez et al., 2000; Konstantinou et al., 2001; Sudo et al., 2002; Cerejeira et al., 2003; Park et al., 2005; Claver et al., 2006; Gómez-Gutierrez et al., 2006). According to literature, the fraction of molinate lost through each of these dissipation modes varies with the conditions of the field studied. However, several reports showed that volatilization is the major mode of molinate dissipation. For example, Soderquist et al. (1977) reported that 75-85 % of molinate is volatized within a week, while Seiber et al. (1986) and Ross and Sava (1986) found that only about 35 % of the applied molinate was dissipated through volatilization. Quayle et al. (2006) estimated that about 30 % of the applied molinate is instantaneously lost to atmosphere. Under field conditions numerous factors may influence the extent of the molinate dissipation processes, explaining the different results reported in literature. For example, the paddy water temperature strongly influences the molinate losses through volatilization (Soderquist at al., 1977; Quayle et al., 2006), while the organic carbon content influences adsorption of this herbicide to soil (Finocchiaro et al., 2005; Park et al., 2005). In fact, the adsorption of molinate onto soil constituents and/or organic matter, as microbial biomass and colloidal suspensions, may reduce the extent of both volatilization (Deuel et al., 1978) and photodegradation in water (Konstantinou et al., 2001). Although small, the fraction of molinate adsorbed onto soil (= 10 %) persists under flooded conditions and decreases in drained fields (Soderquist at al., 1977; Ross and Sava, 1986), probably due to photodegradation (Konstantinou et al., 2001), or microbial degradation (Imai and Kuwatsuka, 1982; Imai and Kuwatsuka, 1988). Given that rice culture uses a flood depth of about 10-20

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cm, a water flow between the rice paddies and the surrounding water streams must occur. Thus, factors as precipitation and withhold period before drainage of floodwater into receiving water streams may influence both molinate leaching and surface run-off. Contamination of receiving waters through surface run-off has been reported in various countries at levels up to 100 µg l-1 of molinate (Julli and Krassoi, 1995; Mabury et al., 1996; Tanabe et al. 1996; Albanis et al., 1998; Tsuda et al., 1998; Hernandez et al., 2000; Sudo et al., 2002; Cerejeira et al., 2003; Claver et al., 2006; Gómez-Gutierrez et al., 2006).

O N

S

Molinate

OH O N

O

S

O O N

N

S

OH

S

Molinate-alcohol

2-Hydroxy-Molinate Molinate-sulfoxide O N

O

O

S

2-Oxo-Molinate

O N

N

S

COOH

O S

Molinate-acid

O

Molinate-sulfone

NH

Hexamethyleneimine

Figure 1. Proposed molinate degradation pathways. Adapted from Imai and Kuwatsuka, 1986c.

Although molinate is considered one of the most recalcitrant thiocarbamates (Nagy et al., 1995), several studies showed that molinate is microbiologically transformed into several partially oxidized products (Klysheva et al., 1980; Golovleva et al., 1981; Zyakun et al., 1983; Imai and Kuwatsuka, 1986a, 1986b, 1986c). Considering the compounds detected in environmental samples, in experimental soils and in molinate degrading microbial cultures, three main putative molinate degradation pathways were proposed (Fig. 1) (Soderquist et al.,

From Planting to Harvest: Environmental Dissipation of the Herbicide Molinate…

81

1977; Thomas and Holt, 1980; Golovleva et al., 1981; Imai and Kuwatsuka, 1986c). Molinate alcohol and molinate acid are formed through the oxidation of molinate ethyl moiety, while hydroxy- and oxo-molinate are derived from azepane ring oxidation. Sulfur oxidation leads to the formation of molinate sulfoxide and sulfone, and it is supposed that these compounds may undergo a thioester bond cleavage with the release of hexamethyleneimine (HMI). It is reported in literature that maximal concentrations of molinate sulfoxide and HMI were detected in water, respectively, 14 and 7 days after molinate application, representing around 7 % and 9 % of the applied herbicide (SANCO, 2003). In soils, maximal concentrations were detected only after 30 days, and did not exceed 2 % and 1 % for molinate sulfoxide and HMI, respectively (SANCO, 2003). The same sulfur- and/or ring- oxidized products are formed both through molinate photolysis (Soderquist et al., 1977; Konstantinou et al., 2001) and detoxification processes of plant (Imai and Kuwatsuka, 1988) and animal cells (Jewell et al., 1999). Other oxidation processes, as, for example, drinking water chlorination, may also lead to molinate transformation into molinate sulfoxide (Cochran et al., 1997). Despite of being considered a moderately toxic compound, the toxicological information of molinate refers its toxicity to some aquatic animals, as the common carp (Tjeerdema and Crosby, 1988) and the adverse reproductive and neurotoxic effects and possible carcinogenicity for mammals (Cochran et al., 1997; SANCO, 2003). Moreover, molinate sulfoxide has been described as being more persistent and, above all, more toxic to animals than molinate (Golovleva et al., 1981; Tjeerdema and Crosby, 1988; Ellis et al., 1998; Jewell and Miller, 1998; Jewell et al., 1999). Thus, considering the extensive oxidation of molinate that occurs in the environment, the contamination of natural waters with this xenobiotic may have more harmful effects than those assumed for the herbicide itself. Table 1. Selected physicochemical properties of molinate.

O Chemical structure

Molecular weight Relative density Appearance Water solubility (pH unbuffered) Solubility in organic solvents (20 °C) Hydrolytic stability (DT50) (pH 5-9) Dissociation constant UV/VIS absorption

N

S

187.3 g mol-1 1.0643 Pale yellow liquid 1100 mg l-1 Miscible (e.g. hexane, ethyl acetate, ethanol) > 1 year No dissociation occurs λ max. ~ 200 nm (no absorption at λ > 290 nm)

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Célia M. Manaia and Olga C. Nunes Table 1. (Continued)

Freezing point Boiling point Vapor pressure (25 °C) Henry’s law constant (25 °C) Organic coefficient (mean Koc) Soil/water distribution coefficient (mean Kd)

< -25 °C 277.5 - 278.5 °C 0.5 Pa 0.687 Pa m3 mol-1 190 dm3 kg-1 2.9

Data from SANCO, 2003; Finocchiaro et al., 2005.

ENVIRONMENTAL DISSIPATION OF MOLINATE IN A PORTUGESE RICE FIELD Factors as climate conditions (e.g. rain, temperature, humidity, winds regime), soil characteristics, and agricultural practices may influence the dissipation of pesticides in the environment, making the ways of dispersion specific for each time of the year and geographic region. Therefore, it was considered relevant to follow the molinate dispersion during its application in a Portuguese rice field (“Quinta do Seminário”), and to compare the data obtained with those reported for other rice producing countries (Castro et al., 2005). “Quinta do Seminário” is a 70 ha commercial rice field, situated in the valley of river Pranto, a tributary of river Mondego, in central Portugal (40°01’N, 8°41’W). The physicochemical properties o f the rice field soil are presented in Table 2. The field studied has a long history of molinate application - Ordram, a granular compound containing 7.5 % (w/w) molinate, has been applied once a year at a rate of 40 to 60 kg ha-1 for more than 30 years. Every year, in late Winter-early Spring (March) the rice paddies of this farm are prepared for new planting. The preparation of the fields starts with the water aeration, which consists on its drainage from soil, pumping up to about 2 m high, flowing out to a reservoir, and re-circulation throughout the field. About one month later (late April-early May), Ordram is applied to flooded paddies (water depth of about 10 cm) and the rice seedlings are transplanted. After a holding period of about 20 days, the floodwater is drained and the herbicide Stam (propanil, 80 % w/w) and the acidifying agent BB5 are applied at rates of 6 kg ha-1 and 250 ml ha-1, respectively. Two days after Stam application paddies are flooded again (depth of around 10 cm) and the herbicide Rancho (bensulfuron-methyl, 0.083 % w/w, mefenacet 1.25 % w/w, and molinate, 3.75 % w/w) is applied at a rate of 60 kg ha-1. After a period of 3-4 weeks, crop treatment is completed with the application of ammonium sulfate. To monitor molinate losses and the possible contamination of surface and underground waters due to Ordram application, water samples were collected from different sites, during Spring 2002 (Castro et al., 2005). Molinate concentration in these samples was determined by solid-phase micro-extraction (SPME) followed by gas chromatography with flame photometric detector (GC-FPD). In paddy water, the highest molinate concentration (1570 µg l-1) was detected in samples

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83

collected on the day of Ordram application. After 9 days, the molinate concentration in paddy waters decreased to 94-125 µg l-1. Similar results on molinate concentration variations were observed in other rice paddy waters in USA (Soderquist et al., 1977; Deuel et al., 1978) and Australia (Quayle et al., 2006), for equivalent periods of rice crop treatment. The readily dissipation of molinate observed is in agreement with several studies which report a half-life of molinate in paddy water in rice fields of 3-10 days (Deuel et al., 1978; Ross and Sava, 1986; Johnson and Lavy, 1995; Quayle et al., 2006). When the rice crop treatment was complete, corresponding to approximately one month and a half after Ordram application, the molinate concentration in floodwater was about 19 µg l-1, a value similar to that found in drained water. Given that tail waters from “Quinta do Seminário” and from other rice fields of river Pranto valley are discharged into the river during crop phytochemical treatment, it was supposed that river water contamination with herbicide could increase during this period. In fact, it was observed that molinate concentration reached values around 16 µg l-1 in river Pranto, 5 km downstream from the studied rice field, one month and a half after Ordram application. This result is in agreement with several studies that report the detection of molinate in natural receiving waters, with the highest values observed for the same period of crop treatment (Tanabe et al. 1996; Albanis et al., 1998; Tsuda et al., 1998; Sudo et al., 2002; Cerejeira et al., 2003; Claver et al., 2006; Gómez-Gutierrez et al., 2006). It should be noted, however, that an accurate comparison of the molinate concentrations in such water environments reported in these studies is hampered by different hydrological regimes, river water volumes and flow rates, or variable distances between the rice paddy fields and the sampling area. These differences explain the wide range (0.021-58 µg l-1) of maximal molinate concentrations reported in river waters worldwide (Portugal, Spain, Greece, Japan, USA) shortly after (about 1 month) the herbicide application on rice fields. However, after this peak of contamination, the decrease throughout summer to achieve undetectable levels during winter seems a general trend (Mabury et al., 1996; Tanabe et al., 1996; Albanis et al., 1998; Sudo et al., 2002; Cerejeira et al., 2003; Claver et al., 2006). Table 2. Physicochemical properties of paddy field soil from “Quinta do Seminário” Soil Texture Sand (%)

67

Silt (%)

19

Clay (%)

14

Physicochemical analysis pH

5.8

Electric conductivity (mΩ cm-1)

0.3

Total organic matter (%)

1.17

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Célia M. Manaia and Olga C. Nunes Table 2. (Continued). Total nitrogen (%)

0.07

Total organic carbon (%)

0.68

Carbon/Nitrogen ratio

9.7

Average values of a total of four soil samples. Analyzed by Fitosoil, Centro de Agroanalisis, San Gines, Murcia, Spain, on January 2000. Adapted from Castro et al., 2005.

Beside molinate surface run-off, also the leaching of this herbicide was investigated after Ordram application in “Quinta do Seminário”. The molinate concentration in a covered well with 4 meters depth, located nearby the water pump, varied between approximately 4 µg l-1 in March, about one year after the last herbicide application (April 2001), and 477 µg l-1 in May, at the time of Ordram application in 2002. As this well had no direct connection with the drainage canals system, the most plausible explanation for the presence of molinate in the well, is the occurrence of leaching from the rice paddy fields. The potential of a pesticide to contaminate ground water depends on its mobility and persistence (Claver et al., 2006). Molinate has a moderate mobility and a light-medium persistence which make it a possible contaminant. In fact, molinate medium average life (persistence) together with its poor adsorption to soil and high water solubility (Table 1) allow its leaching (Wauchope et al., 1992; Claver et al., 2006). Confirming the potential of molinate to contaminate ground waters, Albanis et al. (1998) reported the detection of molinate in wells located in a rice-fields area in concentrations below 0.08 µg l-1. Also in the present study, the detection of molinate in the drainage water and in the well of “Quinta do Seminário”, one year after its application, is indicative of its persistence in the environment. Such persistence is possibly increased when the herbicide reaches sites where major dissipation modes (volatilization, photo-, biodegradation) are insignificant or inexistent. In this field study was possible to conclude that the utilization of molinate in weeds control led to the contamination of surface and underground water above the legally recommended upper limit for this herbicide (2 µg l-1, Dec-Lei 261/2003). The establishment of treatment methodologies to reduce the degree of environmental contamination with molinate was another objective of this research.

MOLINATE CLEAN-UP METHODOLOGIES Bioremediation has been described as a high efficient and low cost method for the removal of toxic compounds from polluted sites. However, such treatment systems require the utilization of organisms able to mineralize the pollutant, or, at least, to transform the parent molecule into non-toxic degradation products. Additionally, the success of a biological treatment depends on the optimization of cell growth and biodegradation conditions at large scale. Therefore, to establish a biological system for molinate decontamination, firstly, it was necessary to collect organisms able to degrade this herbicide without accumulation of toxic transformation products, as molinate sulfoxide. The use of organisms able to mineralize the molinate seemed the ideal solution, although it was apparent, from previous literature reports,

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85

that organisms with such capacity could be rare in nature. Previous reports referred the microbiological degradation of molinate has being a strictly co-metabolic process, leading to the accumulation of the toxic oxidized metabolites (Zyakun et al., 1983; Imai and Kuwatsuka, 1986a, 1986b, 1986c; Molinari et al., 1992). Several bacterial and fungi strains, isolated from paddy water or soil, are able, in mixed or axenic cultures, to degrade, but not to mineralize, the molinate (Zyakun et al., 1983; Imai and Kuwatsuka, 1982, 1986a, 1986b, 1986c). Among these, strain Fusarium sp. F1 presented the highest mineralization extent, releasing about 25 % of the initial 14C-molinate (40 mg l-1) as 14CO2 with the simultaneous accumulation of ring oxidized products after 10 days of incubation in complex medium (Imai and Kuwatsuka, 1986c). In view of these results, the isolation of organisms able to efficiently mineralize the molinate was a priority in this study.

Isolation and characterization of microorganims able to mineralize molinate A mixed culture (EC1) that simultaneously was able to degrade molinate and to deplete the dissolved organic carbon, i.e. to mineralize the herbicide, was obtained through enrichment of a mixture of soil and water collected nearby a molinate producing industry, in mineral medium with 400 mg l-1 molinate. In this enrichment culture (EC1) molinate was the sole source of carbon and energy (Barreiros et al., 2003) while ammonium sulfate and potassium nitrate were additional nitrogen sources. In order to obtain organisms both able of use molinate as the single nutrient source (energy, carbon and nitrogen) and to mineralize higher concentrations of the herbicide, the culture EC1 was further enriched in mineral medium with 750 mg l-1 molinate. With this procedure a second molinate mineralizing enrichment culture, named EC2, was established. The use of a higher molinate concentration and the absence of other nitrogen sources reduced the 45 cultivable bacterial strains present in enrichment culture EC1 to only 5 in enrichment culture EC2. The five isolates of enrichment culture EC2, designated strains ON1, ON2, ON3, ON4 and ON5, were characterized for their fatty acid methyl esters (MIDI system) and nutritional (Biolog Inc. Microbial Identification System) profiles, and their phylogenetic affiliation was ascertained through the 16S rRNA gene sequence analysis. The combination of these three methodologies allowed the identification of four of these isolates as members of genera Pseudomonas (strains ON1 and ON3), Stenotrophomonas (strain ON2) and Achromobacter (strain ON5). Strain ON4, was a Gram positive actinobacterium that, at the time of its isolation and characterization, could not be affiliated to any validly described species or genus, being the species Curtobacterium flaccumfaciens pv. flaccumfaciens LMG 3645T its closest phylogenetic neighbor (Table 3) (Barreiros et al., 2003). The further phenotypic and phylogenetic characterization of isolate ON4 led to the proposal of a new genus and new species name, Gulosibacter molinativorax, of the family Microbacteriaceae (Fig. 2) (Manaia et al., 2004). In axenic culture, none of the Gram negative isolates (ON1, ON2, ON3, ON5) was able to grow or to degrade molinate as carbon and/or nitrogen source, even in the presence of vitamins or other organic growth factors. In contrast, G. molinativorax ON4T was able to grow and to degrade molinate as the single source of energy, carbon and nitrogen. In spite of being able to degrade the molinate alone, it was observed that herbicide concentrations above 375 mg l-1, corresponding to half of the herbicide concentration used by enrichment culture EC2, inhibited the growth of strain ON4T. Nevertheless, even under growth inhibitory

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Célia M. Manaia and Olga C. Nunes

conditions, G. molinativorax ON4T was still capable of molinate transformation. However, while enrichment culture EC2 was able to mineralize molinate, under those conditions G. molinativorax ON4T could only transform part of the herbicide. When cultured in medium with 750 mg l-1 of molinate, culture EC2 reduced the herbicide to concentrations below the detection limit (HPLC-UV, 0.9 mg l-1) and simultaneous depletion of the dissolved organic carbon (DOC), whereas G. molinativorax ON4T promoted reductions of only 40 % of initial molinate, and led to DOC content accumulation. These results indicated that besides G. molinativorax ON4T, other microorganism(s) present in enrichment culture EC2 were necessary to achieve an efficient molinate mineralization. In fact, when equal proportions of each of five isolates - the four Gram negative (ON1, ON2, ON3, and ON5) and G. molinativorax ON4T were mixed, it was observed that this prepared defined mixed culture (named DC) was able to grow and to mineralize molinate. This observation confirmed that the co-operation between the different enrichment culture (EC2) members was necessary to achieve molinate mineralization. Moreover, the similarity of the kinetic parameters (specific growth and molinate degradation rates) observed for mixed culture DC and for enrichment culture EC2 suggested that the five culturable isolates might be involved in molinate mineralization in enrichment culture EC2 (Barreiros et al., 2003). In a next stage of this study, it was investigated the role of the different organisms in molinate mineralization. Agrococcus jenensis DSM 9580T (X92492)

63

Microbacterium lacticum DSM 20427T (X77441) Okibacterium fritillariae DSM 15271T (AM410675) Plantibacter flavus P 297/02T (AJ310417)

84

Leifsonia aquatica JCM 1368T (D45057) Agromyces ramosus DSM 43045T (X77447) Agreia bicolorata DSM 14575T (AM410672)

99

Subtercola boreus DSM 13056T (AM410674) Rhodoglobus vestalii LV3T (AJ459101)

74 100

Salinibacterium amurskyense KMM 3928T (AF539699) Leucobacter komagatae JCM 9414T (D45063)

Clavibacter michiganensis DSM 46364T (X77435) Cryobacterium psychrophilum DSM 4854T (AM410676) Rathayibacter rathayi DSM

7485T(X77439)

Mycetocola saprophilus DSM 15178T (AM410677) Frigoribacterium faeni DSM 10309T (AM410686) Microcella putealis CV2T (AJ717388)

80 100

Yonghaparkia alkaliphila KSL-133T (DQ256088)

Curtobacterium citreum DSM 20528T (X77436) Pseudoclavibacter helvolus DSM 20419T (X77440) 94

Gulosibacter molinativorax ON4T (AJ306835)

0.005

Figure 2. Phylogenetic tree based on 16S rRNA gene sequences showing the relationship between Gulosibacter molinativorax ON4T and the type strains of the type species of the family Microbacteriaceae. Bootstrap values were generated from 1000 re-samplings, only values greater than 60 % are shown. Bar, 1 substitution per 200 nucleotide positions.

From Planting to Harvest: Environmental Dissipation of the Herbicide Molinate…

87

Table 3. Closest validly described species of the isolates from enrichment culture EC2. Isolate

16S rRNA gene sequence analysis (% identity)

ON1

Pseudomonas chlororaphis IFO 3904T (99.8)

ON2

Stenotrophomonas maltophilia LMG 958T (99.3)

ON3

Pseudomonas nitroreducens IAM 1439T (99.8)

ON4

Curtobacterium flaccumfaciens pv. flaccumfaciens LMG 3645T (95.3)

ON5

Achromobacter xylosoxidans spp. denitrificans ATCC 15173T (99.6)

Adapted from Barreiros et al., 2003.

Role of the bacterial isolates in molinate mineralization One of the first clues to follow molinate mineralization by mixed culture DC was the understanding of the involvement of G. molinativorax ON4T, known as capable of molinate transformation. When using molinate as the only source of energy, carbon and nitrogen (Fig. 3a), G. molinativorax ON4T was able to degrade the herbicide with the concomitant accumulation of ethanethiol and diethyl disulfide (Fig. 3b). The composition of the headspace of G. molinativorax ON4T cultures indicated that, in mixed culture DC, this organism is the first to attack molinate, breaking the thioester bond of the thiocarbamate with the release of ethanethiol (ethyl moiety) and possibly N-carboxy hexamethyleneimine (azepane ring moiety). However, further attempts to detect the azepane ring of molinate or its by-products by HPLC-UV, GC-MS and SPME-GC-FID analysis of head space or of organic extracts of axenic cultures of G. molinativorax ON4T were vain, suggesting that this organism might readily use the herbicide ring moiety as source of energy, carbon and nitrogen. Such assumption was consistent with the ability of strain G. molinativorax ON4T to grow at expenses of molinate. The observation that ethanethiol and diethyl disulfide (formed spontaneously from ethanethiol, Fig. 3a) were neither accumulated in the headspace of mixed culture DC, nor in a culture of G. molinativorax ON4T with a mixed culture of the Gram negative isolates incubated in a vial placed in the headspace, indicated that the Gram negative isolates were responsible for the consumption of these sulfur compounds in mixed culture DC. Moreover, the presence of the Gram negative isolates in the headspace of a G. molinativorax ON4T culture restored the growth of the Gram positive isolate in medium with 750 mg l-1 molinate, confirming that these sulfur compounds were toxic to G. molinativorax ON4T when present at concentrations above 2 mM (Barreiros et al., 2003). These results revealed that molinate degradation by mixed culture DC involves different kinds of interaction among its members, namely metabolic and detoxifying association.

88

Célia M. Manaia and Olga C. Nunes Retention time (min) 1600

H3C

1200

*

O N

S

800

S

CH3

A

S S

400

H3C 1600 1200

H3C

HS

CH3

S

S

CH3

B O

800 400

H3 C

S

* S

CH3

1600 1200

C

800 400

1600 1200

D

800 400

Retention Index Figure 3. Chromatograms obtained by SPME/GC/FID analysis of headspace of different cultures in mineral medium with 187 mg l-1 molinate. Uninoculated medium after 7 days of incubation (A). Culture of isolate ON4 after 7 days of incubation (B). Culture of isolate ON4 with a mixed culture of the Gram negative isolates incubated in a vial placed in the headspace, after 7 days of incubation (C). Culture of defined mixed culture DC after 7 days of incubation (D). (*) Tentatively identified by comparing the Kovats indices and the mass spectra present in the NIST 98 MS Library Database. Retention time of the compounds: ethanethiol, 3.2 min; diethyl disulphide, 24 min; diethyl trisulphide, 44 min; S,S-diethyl ester carbonodithioic acid, 45 min; and molinate, 80 min. (Reprinted with permission from L. Barreiros, B. Nogales, C. M. Manaia, A. C. Silva Ferreira, D. H. Pieper, M. A. Reis, and O. C. Nunes, Environmental Microbiology, 2003, 5: 944-953. © Society for Applied Microbiology and Blackwell Publishing Ltd.).

Evaluation of the feasibility of a water bioremediation treatment system with mixed culture DC The physiological and metabolic properties of the mixed culture DC make it a valuable tool for the implementation of bioremediation processes 2003). Besides being about seven times more tolerant to molinate than described in literature (Klyseva et al., 1980; Golovleva et al., 1981; Imai

described above (Barreiros et al., other organisms and Kuwatsuka,

From Planting to Harvest: Environmental Dissipation of the Herbicide Molinate…

89

1982; 1986a, 1986b, 1986c; Carrasco et al., 1992; Daffonchio et al., 1999), this culture uses the herbicide as the single source of carbon, nitrogen and energy and mineralizes the herbicide without the accumulation of dead-end products, including the toxic oxidized derivatives of molinate (e.g. molinate sulfoxide). Thus, having in mind the use of this culture for bioremediation purposes, it was decided to study the effect of several factors and environmental perturbations on the performance of the degradative activity of mixed culture DC. Major factors affecting microbial activity and, hence, the degrading capacity (namely rate and extent of degradation) are the temperature, the pH, and the availability and type of nutrients and electron acceptors. Therefore, the effect of these factors on mixed culture DC performance was studied with two main objectives – establish the adequate conditions for the implementation of a bioremediation treatment system and assess the vulnerability of the degradative community to environmental perturbations. The effect of operating conditions on molinate mineralization by mixed culture DC was studied in batch cultures grown with controlled temperature, pH, stirring and aeration rate (Correia et al., 2006). Molinate mineralization efficiency under each operating condition assayed was ascertained on the basis of the determined specific cell growth (µ), molinate (rs) and DOC (rs’) degradation rates, cell growth yield (Y), percentage of DOC consumed and accumulation of molinate degradation products. For each factor studied, it was tested a range of operating conditions, similar to those found in natural environments or in biological treatment systems. For example, the range of temperatures assayed was 15-35 °C, which corresponds to Winter-Summer average temperatures in river and paddy waters. The salinity tested simulated the tide effect on the receiving natural water streams (rice fields areas may be located at 5-10 km from the river mouth). The addition of complex organic nutrients simulated the occurrence of an organic overload, for example due to a discharge or algal bloom and which, being preferred nutrients, could hamper biodegradation. These assays revealed the high stability of the physiological and metabolic properties of mixed culture DC, which maintained high degradative capacity under a different array of conditions. Under aerobic conditions, mixed culture DC grew and degraded 187 mg l-1 of molinate, that was used as the single source of energy, carbon and nitrogen, in a wide range of operating conditions - temperature (15-35 °C), pH (6.3-8.0), and salinity (0.04-1.0 % NaCl, w/v). In all these conditions molinate mineralization was evidenced by a direct association between molinate removal and DOC depletion. Among these operating conditions, temperature was the factor that most affected the kinetic parameters of mixed culture DC. The highest specific cell growth and degradation rates (0.11 h-1 and 0.180 gmolinate g-1cell dry weight h-1) and the lowest (0.01 h-1 and 0.027 gmolinate g-1cell dry -1 weight h ) were observed in cultures grown at 35 and 15 °C, respectively. The lowest extent of molinate mineralization was also registered in cultures grown at 15 °C, with 55 % of initial DOC content reduction. In contrast, the highest values of initial DOC depletion ranged between 77 and 85 %, and were observed in cultures grown at 30 or 35 °C, with aeration rates of 0.1 or 0.2 volume of air per volume of reactor per minute (corresponding to oxygen concentrations of 3 and 5 mg l-1, respectively) (Table 4). The presence of alternative and readily consumable organic nutrients, as yeast and meat extracts in concentrations up to 1 g l-1 each, did not affect molinate degradation by mixed culture DC. In fact, the results obtained suggested that the addition of organic nutrients may even improve molinate degradation, as these nutrients favored the specific cell growth and DOC depletion rates and DOC removal, when compared with cultures without organic supplement (Table 4).

90

Célia M. Manaia and Olga C. Nunes

Table 4. Specific growth (μ), molinate (rs) and dissolved organic carbon (rs’) degradation rates, precent of dissolved organic carbon reduction and cell growth yield (Y, Y’) of culture DC grown in mineral medium with 187 mg l-1 of molinate under different batch operating conditions. Light gray shadowing indicates default operating conditions.

Temperature (°C)

rs

µ

% DOC -1

(g molinate g cell

(h-1)

dry wt h-1)

reduction

Y (g cell dry wt g-1 molinate)

15

0.010

0.027

55

0.38

20

0.022

0.046

73

0.49

30

0.058

0.110

77

0.51

35

0.110

0.180

84

0.60

µ

rs

pH

% DOC

(h-1)

-1

(g molinate g cell dry wt h-1)

reduction

Y (g cell dry wt g-1 molinate)

6.3

0.057

0.130

66

0.43

7.2

0.058

0.110

77

0.51

8.0

0.045

0.140

74

0.32

rs

% DOC

Aeration (vvm)

µ (h-1)

0.1 (3 mg O2 l-1) -1

0.2 (5 mg O2 l ) Salinity (% NaCl, w/v)

-1

(g molinate g cell -1

dry wt h )

reduction

Y (g cell dry wt g-1 molinate)

0.064

0.090

85

0.70

0.058

0.110

77

0.51

rs

% DOC

µ (h-1)

-1

(g molinate g cell dry wt h-1)

reduction

Y (g cell dry wt g-1 molinate)

0.04

0.058

0.110

77

0.51

1

0.038

0.100

62

0.38

r’s

% DOC

Additional nutrients

µ (h-1)

-1

(g DOC g cell dry wt h-1)

reduction

Y’ (g cell dry wt g-1 DOC)

Yeast, meat extracts*

0.240

0.230

80

1.0

Control (molinate)

0.058

0.060

77

0.9

* 1 g l-1 each. Adapted from Correia et al., 2006.

The HPLC-UV analysis of ethyl acetate extracts of culture liquid phases showed that trace amounts of molinate and of its oxidized derivates 2-oxo-molinate and molinate

From Planting to Harvest: Environmental Dissipation of the Herbicide Molinate…

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sulfoxide were present at the end of the incubation periods. The concentration of 2-oxomolinate did not exceed 0.088 mg l-1 (30 °C, pH 6.3, 0.2 vvm, 0.04 % NaCl), and molinate sulfoxide was only detected (0.029 mg l-1) in cultures supplemented with 1 g l-1 of yeast and meat extracts, suggesting that these compounds were not major metabolic degradation products. Given that also the molinate concentrations found were low, ranging between below the detection limit (< 0.025 mg l-1) (e.g. 30 °C, pH 7.2, 0.2 vvm, 0.04 % NaCl) and 10 mg l -1 (30 °C, pH 7.2, 0.2 vvm, 1 % NaCl), the values of residual DOC concentration detected at the end of the incubation period (35-40 mg l-1) were attributed to organic compounds released through cell lysis. These results showed that mixed culture DC was capable of efficient aerobic molinate mineralization under a wide range of operating conditions, with cell growth yields ranging between 0.038 gcell dry weight g-1molinate for cultures grown with 187 mg l-1 molinate as single nutrient and stressful operating conditions (15 °C, pH 7.2, 0.2 vvm, 0.04 % NaCl, or 30 °C, pH 7.2, 0.2 vvm,1 % NaCl) and 1 gcell dry weight g-1DOC for cultures grown in the same molinate containing media supplemented with alternative nutrients (30 °C, pH 7.2, 0.2 vvm, and 0.04 % NaCl, 1 g l-1 yeast and meat extracts, each). When compared with aerobic processes, bioremediation systems conducted under anoxic conditions are usually preferable because of the lower cell growth yields obtained under microaerobic to anaerobic conditions. Given that most of the mixed culture DC members are able to reduce nitrate, the growth and mineralization of molinate under anoxic conditions, with nitrate as the final electron acceptor was studied. However, with culture DC, it was observed that among the conditions tested, aeration was the factor that most affected molinate mineralization. Under anoxic conditions, growth did not occur and molinate was partially degraded (90 %), with only 30 % of the initial DOC content being removed after 3.5 days of incubation. These results showed that under anoxic conditions the herbicide was not mineralized, indicating that molinate mineralization by mixed culture DC is highly dependent on oxygen. Similar findings were reported for other molinate degrading microorganisms (Thomas and Holt, 1980; Imai and Kuwatsuka, 1982).

Molinate degradation by mixed culture DC in a continuous reactor Mixed culture DC was grown in a continuous stirring tank reactor (CSTR) operated at different feeding molinate concentrations (187 to 560 mg l-1) with a hydraulic residence time (HRT) of 3.5 days. Molinate constituted the single organic nutrient, and other operating conditions, as temperature, pH, salinity and aeration were set at values previously known as non-limiting (30 °C, pH 7.2, 0.2 vvm, and 0.04 % NaCl) (Correia et al., 2006). Under these conditions, mixed culture DC was able to reduce molinate concentrations from 520 mg l-1 to 3 mg l-1, with a mean degradation rate of 0.07 gmolinate g-1cell dry weight h-1, similar to the rates registered in batch reactors operated with the same conditions (0.11 gmolinate g-1cell dry weight h-1) (Fig. 4). When compared with other molinate degraders described in the literature, mixed culture DC is considerably more efficient. Carrasco et al. (1992) reported that a mixed microbial culture was able to remove 97 % of molinate in a CSTR fed with 50 mg l-1 molinate as the carbon source and ammonia as nitrogen source, at a hydraulic retention time (HTR) of 23 days. Mixed culture DC showed capacity to use higher molinate concentrations, to use the

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herbicide as the single source of energy, carbon and nitrogen, and to remove more than 99% of the herbicide with a HTR of 3.5 days.

Figure 4. Evolution of biomass and molinate in a CSTR operated with a HTR of 83 h, fed with different molinate initial concentrations (1-3 mM). Symbols: ---▲---, Biomass concentration; ---■---, Molinate concentration in the effluent; ▬▬▬, Molinate concentration in the feed. (Reprinted with permission from P. Correia, R. A. R. Boaventura, M. A. Reis, and O. C. Nunes, Water Research, 2006, 40: 331-340. © Elsevier Ltd).

Removal of molinate from real-world waters by mixed culture DC Mixed culture DC showed to be highly efficient on molinate mineralization in synthetic mineral medium with the herbicide as single source of energy, carbon and nitrogen or even in the presence of alternative energetically favourable organic nutrients. Having in mind the potential application of this mixed culture in a biological wastewater treatment system, it was considered relevant to study its behaviour in real-world contaminated waters. Aspects as the absence of mineral nutrients in appropriate concentrations, the presence of other naturally occurring organic compounds, including toxic substances, and the effect of the natural microbiota could hamper molinate mineralization. The growth and degradative performance of mixed culture DC was compared in naturally contaminated waters and in synthetic medium. Mixed culture DC was grown, aerobically (0.2 vvm) at 30°C, in river Pranto water, collected in a rice field area soon after molinate application (June 2002). Selected physicochemical properties of river Pranto water are shown in Table 5.

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Table 5. Physicochemical properties of river Pranto water.

pH

8.0 -1

Electric conductivity (µS cm )

1090 -1

Dissolved organic carbon (mg l ) -1

Molinate concentration (µg l )

1.3 16

Adapted from Correia et al., 2006.

The river water was spiked with 187 mg l-1 molinate and no other nutrient was added. Under these conditions, mixed culture DC grew and mineralized molinate as efficiently (µ= 0.042 h-1; rs= 0.096 gmolinate g-1cell dry weight h-1; % DOC depletion = 70; Y= 0.44 gcell dry weight g1 molinate) as when grown in synthetic medium with equivalent operating conditions (µ= 0.045 h-1; rs= 0.140 gmolinate g-1cell dry weight h-1; % DOC depletion = 74; Y= 0.32 gcell dry weight g-1molinate) (Fig. 5). Beside residual molinate (1.68 mg l-1), only low concentrations of 2-oxo-molinate (0.13 mg l-1) were detected in the culture at the end of the incubation period (Correia et al., 2006). The previous assays, whether conducted in synthetic medium or in real-world waters were carried out in the presence of high molinate concentrations, which hardly can be found, even in severely contaminated water streams. In real-world waters, the highest molinate concentrations found were around 1-2 mg l-1 in the paddy field water on the day of the herbicide application or on the following day (e.g. Soderquist et al., 1977; Castro et al., 2005; Quayle et al., 2006). Thus, despite of being able to efficiently mineralize high concentrations of molinate in real-world waters, a reduced bioavailability of the herbicide in naturally contaminated waters might represent an important limitation for mixed culture DC performance in a biological treatment system. To test the ability of mixed culture DC to remove molinate when this herbicide was at low concentrations, naturally contaminated waters were inoculated with this culture and the molinate concentration was monitored (Castro et al., 2005). Mixed culture DC was grown aerobically in paddy field waters from “Quinta do Seminário” and in water from a well located in the same farm, containing between 11 and 1570 µg l-1 molinate. After three days of incubation at 30 °C, the molinate concentration was bellow the legal recommended upper limit (2 µg l-1, Decreto-Lei 231/2003) in three of these cultures (Table 6). However, in the culture with the highest initial molinate concentration, the final herbicide concentration was slightly higher (2.2 µg l-1), suggesting that a limiting factor, for example a mineral nutrient, may halt molinate mineralization by mixed culture DC. Although these results showed that mixed culture DC is able to eliminate molinate from contaminated real-world waters, the implementation of a bio-treatment process of rice culture effluents may present some practical and technical difficulties. The use of a non-invasive method, capable of rapid cleaning-up of the contaminated effluents on site might represent a valuable complement for the bio-treatment that was being planned. In this respect, the adsorption of molinate to a substrate suitable for further regeneration seemed an appropriate solution.

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Figure 5. Cell growth, molinate degradation and DOC consumption by culture DC grown in water from the river Pranto spiked with 1 mM molinate (T = 30 °C and aeration rate = 0.2 vvm). Symbols: ---♦---, Biomass concentration; ---■---, Molinate concentration in culture supernatant; ---▲---, DOC concentration in culture μ.(t −t0 ) supernatant; □, Molinate concentration in uninoculated control; ______, Fitting of equation X = X 0 .e

rs . X 0

(e μ .(t −t 0 ) − 1) for molinate (Reprinted with permission from P. μ Correia, R. A. R. Boaventura, M. A. Reis, and O. C. Nunes, Water Research, 2006, 40: 331-340. © Elsevier Ltd).

for biomass and of equation S = S 0 −

Table 6. Molinate concentration (μg l -1) after treatment of real-world contaminated waters with mixed culture DC. Days after Ordram application

Un-inoculated control

Mixed culture DC

Well

*

476.0

0.6

Paddy field 1

0

1550.0

2.2

Paddy field 2

44

193.7

< 0.48

Paddy field 3

46

10.5

< 0.48

Water sample

* Not applicable; the sample was collected after 9 days of the first molinate application in the commercial rice farm. Analytical method (SPME/GC-FPD) detection limit of 0.48 µg l-1. Adapted from Castro et al., 2005.

Feasibility study of a combination of adsorption / bio-regeneration for the removal of molinate from contaminated waters Adsorption is a physical treatment process widely used for the removal of pollutants from contaminated waters. Such an approach can promote simultaneously the pollutant removal

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and its concentration on the adsorbent. However, this procedure cannot be regarded as a true treatment process, as the contaminant is simply transferred from the water to the adsorbent without any effective elimination of the pollutant. Complementary processes are thus required, namely those involving the regeneration of the adsorbent with the simultaneous elimination of the pollutant. Among the different regeneration processes available, bioregeneration, i.e. the biological degradation of the adsorbate with the restitution of the adsorbent properties (Aktas and Çeçen, 2007), has been described as an effective option to costly traditional physicochemical processes, as thermal desorption. Moreover, the combination of adsorption and biological degradation has the particular advantage of increasing the bioavailability of the pollutant, which may be a limiting factor for an extensive bio-treatment of real word contaminated waters. Some systems combine the processes of adsorption and biodegradation in a single step, with advantages in comparison to traditional suspended biomass biological reactors. One of the most commonly used of such systems, is constituted by activated carbon and immobilized cells, and receives the name of BAC (biological-activated-carbon). For example, Feakin et al. (1995), using a BAC system, where the solid matrix was simultaneously the adsorbent of the pollutant and an adherence surface for a degrading biofilm, reported the successful removal of toxic herbicides (atrazine and simazine) from potable water. In spite of the good results that this technology can offer, the potential contamination of the effluent with microorganisms may be regarded as a serious disadvantage. This problem can be avoided using a two-step methodology, consisting on the removal and concentration of the pollutant through adsorption, followed by a subsequent biodegradation stage, with the simultaneous regeneration of the adsorbent. Another advantage of this procedure is that the adsorption and bio-regeneration processes may occur on different places and different times. This aspect can be particularly useful for the agriculture effluents treatment. To study the feasibility of such approach in molinate contaminated waters treatment with mixed culture DC, it was necessary to characterize the adsorption of the herbicide onto adequate materials and then assess the potential of their bio-regeneration. After a previous selection, further studies were conducted with resin Amberlite XAD-4 and commercial granular activated carbons (Silva et al., 2004; Coelho et al., 2006). Given the high porosity and surface area of these materials, both were efficient molinate adsorbents, being able to remove molinate from real contaminated waters to values below 0.48 µg l-1 (Castro et al., 2005). The Freundlich equation fitted the molinate adsorption equilibrium isotherm experimental data for resin Amberlite XAD-4 (Silva et al., 2004) while Langmuir fitted the data for commercial activated carbon (Norit GAC 1240 PLUS) (Coelho et al., 2006). The calculated parameters are shown in Table 7. When different conditions of temperature (10-30 °C), electrical conductivity (0-489 mS m-1) and pH (5-9) were tested, no significant differences on the molinate adsorption onto the synthetic resin were observed. For the bioregeneration of the adsorbent it is necessary the prior desorption of the pollutant into the bulk liquid (Aktas and Çeçen, 2007), which only can take place if the adsorption is reversible. At this stage, the microbiota may play a central role, as through the biodegradation process it promotes the reduction of the pollutant concentration in the bulk liquid and thus, leads to its continuous desorption. Given that adsorption of molinate onto resin XAD-4 and activated carbon was reversible, mixed culture DC was able to use the adsorbed molinate as the only source of carbon, nitrogen and energy, promoting the adsorbents bio-regeneration (Table 8). However, the adsorbents bio-regeneration was not complete, because the extent of molinate

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degradation was limited by the equilibrium concentrations reached. Presumably, at least part of the adsorbed molinate occupies adsorbent micropores, where it is strongly adsorbed and thus, unavailable for biodegradation. Therefore, adsorbents with larger pores (mainly meso and macropores) seem to be more adequate for this application. 1

Table 7. Parameters of the Freundlich ( qe = a Ce n ) and Langmuir ( qe = qm

K LCe ) 1 + K LCe

models. The fitted experimental data were obtained with different adsorbent masses and solutions of molinate with different concentrations, prepared in mineral medium B (Barreiros et al., 2003) at 30°C. Adsorbent a n qm KL (mg g-1) (l mg-1)

Resin Amberlite XAD-4

93.64

3.01

NA

NA

Activated Carbon Norit

NA

NA

370

1.59

NA, not applicable. Adapted from Silva et al., 2004 and Coelho et al., 2006.

Table 8. Molinate solid phase concentrations after the adsorption step and after bioregeneration with mixed culture DC, grown with the loaded molinate as the single organic nutrient at 30°C, 120 rpm. Adsorbent Molinate concentration (mgmolinate g-1dry adsorbent) After adsorption step After bio-regeneration step

Resin Amberlite XAD-4

Activated Carbon Norit

150

60

119

60

67

67

340

127

Adapted from Silva et al., 2004 and Coelho et al., 2006.

CONCLUSION One of the major disadvantages of the use of pesticides is the contamination of air, waters, soils and even the trophic chain with xenobiotic compounds or their partial degradation products. Due to its anthropogenic nature, most of these pesticides are unrecognizable for natural microbiota and thus recalcitrant to biodegradation. Under these circumstances, physical processes as adsorption or (photo)chemical degradation may contribute to reduce the pesticide contamination, although may be insufficient to promote a complete and readily decontamination. The herbicide molinate, used worldwide to control rice paddy weeds, is an example of such xenobiotics. Agriculture effluents prefigure a particular situation of water contamination, as the non-confinement of the waters puts serious difficulties to the implementation of treatment processes. Rice paddy fields, in this respect,

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are an interesting exception as the water confinement allows the use of some treatment strategies. Due to their extreme adaptability to environmental conditions, microorganisms may, after a long term exposure to a specific xenobiotic, develop new degradative properties. The mixed culture described along this chapter seems to be one of such examples - it was isolated from an industrial effluent of a molinate producing plant. This mixed culture revealed an exceptional capacity to mineralize molinate, at high rates, under a diverse set of physicochemical conditions. Biodegradation was equally efficient when molinate was close to its solubility limit or in trace levels, in synthetic medium or in river and paddy water, demonstrating the potential of this culture to be used as a bioremediation tool. In the mixed culture, constituted by 5 members, the most important role, the molinate breakdown, is promoted by an actinobacterium G. molinativorax ON4T, through a new metabolic pathway. Although it is admissible that other molinate degrading organisms with similar properties may exist in nature, the uniqueness of this degrading organism is patent both on the fact that strain ON4T is the only member of this new genus and on the novelty of the molinate degradation process this organism is able to perform. In spite of the confinement of the rice culture effluents, the direct use of a biodegradative culture to decontaminate the paddy waters would not constitute an adequate bio-treatment process. Instead, the removal of the molinate from these waters using a rapid and non-invasive method and its subsequent biodegradation seems more appropriate. One of such non-invasive methods to remove molinate from rice paddy waters is the adsorption. With this objective in mind, the adsorbents Amberlite XAD-4 and activated carbon were tested with success. After molinate adsorption, these adsorbents can be regenerated with the aid of mixed culture DC which, promoting the continuous degradation of molinate in the bulk liquid, leads to the continuous desorption of molinate. As adsorption and desorption/biodegradation can be carried on at different places, the first stage must be implemented in the rice paddy fields whereas the second may be conducted in a different location, where a reactor can be installed. This study briefly summarizes the contributions that different scientific areas may bring to solve a real-world problem. From the exploitation of the microbial diversity and physiological properties to the implementation of engineering solutions, the study presented evidences the multidisciplinary character of the environmental sciences.

REFERENCES Albanis, T. A., Hela, D. G., Sakellarides, T. M. & Konstantinou, I. K. (1998). Monitoring of pesticide residues and their metabolites in surface and underground waters of Imathia (N. Greece) by means of solid-phase extraction disks and gas chromatography. Journal of Chromatography A, 823, 59-71. Aktas, O. & Çençen, F. (2007). Bioregeneration of activated carbon: A review. International Biodeterioration & Biodegradation, 59, 257-272. Barreiros, L., Nogales, B., Manaia, C. M., Silva-Ferreira, A. C., Pieper, D.H., Reis, M.A. & Nunes, O.C., (2003). A novel pathway for mineralization of the thiocarbamate herbicide molinate by a defined bacterial mixed culture. Environmental Microbiology, 5, 944-953.

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Carrasco, J. M., Sabater, C., Alonso, J. L., Gonzalez, J., Botella, S., Amoros, I., Ibanez, M. J., Boira, H. & Ferrer, J. (1992). Molinate decontamination processes in effluent water from rice fields. Science Total Environment, 123/124, 219-232. Castro, M., Manaia, C. M., Silva Ferreira, A. C. & Nunes, O. C. (2005). A case study of molinate application in a Portuguese rice field: herbicide dissipation and proposal of a clean-up methodology. Chemosphere, 59, 1059-1065. Cerejeira M. J., Viana, P., Batista, S., Pereira, T., Silva, E., Valerio, M. J., Silva, A., Ferreira, M. & Silva-Fernandes, A.M. (2003). Pesticides in Portuguese surface and ground waters. Water Research, 37, 1055-1063. Claver, A., Ormad, P., Rodriguez, L. & Ovelleiro, J. L. (2006). Study of the presence of pesticides in surface waters in the Ebro river basin (Spain). Chemosphere, 64, 1437-1443. Cochran, R. C., Formoli, T. A., Pfeifer, K. F. & Aldous, C. N. (1997). Characterization of risks associated with the use of molinate. Regulatory Toxicology and Pharmacology, 25, 146-157. Coelho, C., Oliveira, A. S., Pereira, M. F. & Nunes O. C. (2006). The influence of activated carbon surface properties on the adsorption of the herbicide molinate and the bioregeneration of the adsorbent. Journal of Hazardous Materials, 16, 138, 343-349. Correia, P., Boaventura, R. A., Reis, M. A. & Nunes O. C. (2006). Effect of operating parameters on molinate biodegradation. Water Research, 40, 331-340. Daffonchio, D., Zanardini, E., Vatta, P. & Sorlini, C. (1999). Cometabolic degradation of thiocarbamate herbicides by Streptomyces sp. strain M2 and effects on the cell metabolism. Annali di Microbiologia ed Enzimologia, 49, 13-22. Decreto-Lei 261/2003, 21 October 2003. Anexo. Objectivos de qualidade para determinadas substâncias perigosas incluídas nas famílias ou grupos de substâncias da lista II do anexxo XIX ao Decreto-Lei nº 236/98, de 1 de Agosto. Diário da República, série I-A, nº 244. Deuel, L. E., Turner, F. T., Brown, K. W. & Price, J. D. (1978). Persistence and factors affecting dissipation of molinate under flooded rice culture. Journal Environmental Quality, 7, 373-377. Ellis, M. K., Richardson, A. G., Foster, J. R., Smith, F. M., Widdowson, P. S., Farnworth, M. J., Moore, R. B., Pitts, M. R. & Wickramaratne G. A. (1998) The reproductive toxicity of molinate and metabolites to the male rat: effects on testoterone and sperm morphology. Toxicology Applied Pharmacology 151, 22-32. Feakin, S. J., Gubbins, B., MacGhee, I., Shaw, L. J. & Burns, R. G. (1995). Inoculation of granular activated carbon with-s-triazine-degrading bacteria for water treatment at pilotscale. Water Research 29, 1681-1688. Finocchiaro, R. Meli, S. M., Cignetti, A. & Gennari, M. (2005). Adsorption of molinate, terbuthylazine, bensulfuron-methyl, and cinosulfuron on different Italian soils. Fresenius Environmental Bulletin, 14, 690-697. Golovleva, L. A., Finkelstein, Z. I., Popovich, N. A. & Skryabin, G. K. (1981). Transformation of ordram by microorganisms. Izvestiia Akademii Nauk SSSR Seriya Biologicheskaya, 3, 348-358. Gomez-Gutierrez, A. I., Jover, E., Bodineau, L., Albaiges, J. & Bayona, J. M. (2006). Organic contaminant loads into the Western Mediterranean Sea: estimate of Ebro River inputs. Chemosphere, 65, 224-236.

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Hernandez, F., Beltran, J., Lopez, F. J. & Gaspar, J. V. (2000). Use of solid-phase microextraction for the quantitative determination of herbicides in soil and water samples. Analytical Chemistry, 72, 2313-2322. Imai, Y. & Kuwatsuka, S., 1982. Degradation of the herbicide molinate in soils. Journal Pesticide Science, 7, 487-497. Imai, Y. & Kuwatsuka, S. (1986a). Characteristics of microflora degrading the herbicide molinate in soil. Journal Pesticide Science, 11, 57-63. Imai, Y. & Kuwatsuka, S. (1986b). The mode of metabolism of the herbicide molinate by four strains of microorganisms isolated from soil. Journal Pesticide Science, 11, 111-117. Imai, Y. & Kuwatsuka, S. (1986c). Metabolic pathways of the herbicide molinate in four strains of isolated soil microorganisms. Journal Pesticide Science, 11, 245-251. Imai, Y. & Kuwatsuka, S. (1988). Residues of the herbicide molinate and its degradation products in pot soil and rice plants. Journal Pesticide Science 13, 247-252. Jewell, W. T., Hess, R. A. & Miller, M. G. (1999). Comparison of human and rat metabolism of molinate in liver microsomes and slices. Drug Metabolism and Disposition. 27, 842847. Jewell, W. T. & Miller, M. G. (1998). Testicular toxicity in the rat: metabolic activation via sulfoxidation. Toxicology Applied Pharmacology 149, 159-166. Jiménez, B., Moltó, J. C., Font, G. & Soriano, J. M. (1999). Evaluation by HPLC-UV of polar pesticides in rice fields. Bulletin of Environmental Contamination & Toxicology, 63, 813820. Johnson, W. G. & Lavy, T. L. (1995). Organic chemicals in the environment. Persistence of carbofuran and molinate in flooded rice culture. Journal of Environmental Quality 24, 487-493. Julli, M. & Krassoi, F. R. (1995). Acute and chronic toxicity of the thiocarbamate herbicide, molinate, to the Cladoceran Moina australiensis Sars. Bulletin Environmental Contamination Toxicology 54, 690-694. Klysheva, A. L., Golovleva, L. A. & Ilyaletdinov, A. N. (1980). Transformation of ordram by microorganisms isolated from soils of Kazakhstan rice paddies. Izvestiia Akademii Nauk SSSR Seriya Biologicheskaya, 4, 29-34. Konstantinou, I. K, Zarkadis, A. K. & Albanis, T. A. (2001). Photodegradation of selected herbicides in various natural waters and soils under environmental conditions. Journal of Environmental Quality, 30, 121-130. Mabury, S. A., Cox, J. S. & Crosby, D. G. (1996). Environmental fate of rice pesticides in California. Reviews of Environmental Contamination & Toxicolog, 147, 71-117. Manaia, C. M., Nogales, B., Weiss, N. & Nunes, O. C. (2004). Gulosibacter molinativorax gen. nov., sp. nov., a molinate degrading bacterium, and classification of “Brevibacterium helvolum” DSM 20419 as Pseudoclavibacter helvolus gen. nov., sp. nov. International Journal Systematic and Evolutionary Microbiology, 54, 783-789. Molinari, G. P., Sorlini, C. Daffonchio, D., Baggi, G. & Ruffo, L. (1992). Activity and evolution of mixed microbial culture degrading molinate. Science Total Environment, 123/124, 309-323. Nagy, I., Schoofs, G., Compernolle, F., Proost, P., Vanderleyden, J. & De Mot, R. (1995). Degradation of the thiocarbamate herbicide EPTC (S-Ethyldiproylcarbamothioate) and biosafening by Rhodococcus sp. strain NI86/21 involve an inducible cytochrome P-450 system and aldehyde dehydrogenase. Journal of Bacteriology, 177, 676-687.

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Park, B. J., Kyung, K. S., Choi, J. H., Im, G. J., Kim, I. S. & Shim, J. H. (2005). Environmental fate of the herbicide molinate in a rice-paddy-soil lysimeter. Bulletin of Environmental Contamination and Toxicology, 75, 937-944. Quayle W. C., Oliver D. P. & Zrna, S. (2006). Field dissipation and environmental hazard assessment of clomazone, molinate, and thiobencarb in Australian rice culture. Journal of Agricultural and Food Chemistry, 54, 7213-7220. Ross, L. J. & Sava, R. J. (1986). Fate of thiobencarb and molinate in rice fields. Journal of Environmental Quaity, 15, 220-224. SANCO/3047/99-Final, 3 June 2003. Review report for the active substance molinate. European Commission. Health and consumer protection directorate-general. Seiber, J. N., McChesney, M. M., Sanders, P. F. & Woodrow, J. E. (1986). Models for assessing the volatilization of herbicides applied to flooded rice fields. Chemosphere, 15, 127-138. Silva, M., Fernandes, A., Manaia, C. M., Mendes, A. & Nunes, O. C. (2004). Preliminary feasibility study for the use of an adsorption/bio-regeneration system for molinate removal from effluents. Water Research, 38, 2677-2684. Soderquist, C. J., Bowers, J. B. & Crosby, D. G. (1977). Dissipation of molinate in a rice field. Journal of Agricultural and Food Chemistry, 25, 940-945. Sudo M., Kunimatsu T. & Okubo T. (2002). Concentration and loading of pesticide residues in Lake Biwa basin (Japan). Water Research, 36, 315-329. Tanabe, A., Mitobe, H., Kawata, K. & Sakai, M. (1996). Monitoring of herbicides in river water by gas chromatography-mass spectrometry and solid-phase extraction. Journal of Chromatography A, 754, 159-168. Thomas, V. M. & Holt, C. L. (1980). The degradation of [14C]molinate in soil under flooded and nonflooded conditions. Journal of Environmental Science and Health Part B, 15, 475-484. Tjeerdema, R. S. & Crosby, D. G. (1988). Disposition, biotransformation, and detoxification of molinate (ordram) in whole blood of the common carp (Cyprinus carpio). Pesticide Biochemistry and Physiology 31, 24-35. Tsuda, T., Kojima, M., Harada, H., Nakajima, A. & Aoki, S. (1998). Pesticides and their oxidation products in water and fish from rivers flowing into lake Biwa. Bulletin of environmental contamination and toxicology, 60, 151-158. Wauchop, R. D., Buttler, T. M., Hornsby, A. G., Augustijn-Beckers, P. W. & Burt, J. P. (1992). The SCS/ARS/CES pesticide properties database for environmental decisionmaking. Reviews of Environmental Contamination & Toxicology, 123, 1-164. Zyakun, A. M., Nefedova, M. Y., Baskunov, B. P. & Finkelstein, Z. I. (1983). The new products of ordram microbial degradation. Izvestiia Akademii Nauk SSSR Seriya Biologicheskaya 1, 126-130.

In: Environmental Biodegradation Research Focus Editor: B. Y. Wang, pp. 101-126

ISBN: 978-1-60021-904-7 © 2007 Nova Science Publishers, Inc.

Chapter 4

NATURAL ATTENUATION OF HIGH CONCENTRATIONS OF ORGANIC POLLUTANTS BY BIODEGRADATION IN SOILS L. Reijnders1 IBED/ECDO University of Amsterdam, Amsterdam, The Netherlands

ABSTRACT In view of its relatively low cost, monitored natural attenuation by biodegradation is increasingly relied upon to clean up pollution of soils caused by landfills, industrial activities and major transport and storage related spills. Current policy tends to aim at reducing soil pollution to levels reflecting tolerable risk for specific recipients within a reasonable time frame. Natural attenuation by biodegradation of high concentrations of major pollutants that preferentially partition to the particle fraction of soils tends to be poor. These pollutants include hazardous hydrocarbons, highly halogenated hydrocarbons and nitro-organics. Relying on natural attenuation of these compounds, when feasible at all, leads to exceeding tolerable risk levels for a long time. Better perspectives exist for natural attenuation by biodegradation of pollutants that partition to a significant extent to the aqueous phase, especially for low molecular weight organic aromatics and chlorinated solvents. Dependent on conditions in the aquifer and the presence of suitable microorganisms, there can be substantial biodegradation of a variety of hydrocarbons, organochlorines and oxygenates. In some cases natural attenuation has been found to bring down high levels of pollution to levels meeting current standards of tolerable risk. Predictions whether in the future levels reflecting tolerable risk can be attained, are uncertain. Uncertainty is especially large in case of expanding plumes. There are uncertainties that beset current modelling to predict future concentrations and there is uncertainty about what in the future will be considered tolerable risk. Even when it is supposed that current standards will be applied indefinitely, uncertainties related to 1

IBED/ECDO University of Amsterdam, Nieuwe Achtergracht 166, 1018 WV Amsterdam, The Netherlands; Tel.: + 31-20-5256206; Fax: + 31-20- 5257431; E-mail: [email protected]

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L. Reijnders present modelling and sampling often prevent certainty that these standards will be met in the future. In practice, natural attenuation is often falling short of attaining promised outcomes. This means that often interventions aimed at enhanced remediation will be necessary to achieve tolerable risk within a reasonable time frame. These may include enhanced biodegradation.

1. INTRODUCTION Industrial activities have led to soil contamination with high concentrations of a variety of organic compounds, which affects the particulate fraction and groundwater. Long ago courts found that such soil contamination can do harm. For instance, in the early 20th century the case of Ballantine and Sons versus Public Service Corporation was heard at US courts, in which it was established that soil pollution by coal tar wastes infringed the right of a brewery to obtain ‘pure water’ from beneath its property (Colten 1991). In 1943 a local court in US city of Lansing (Michigan) banned the disposal into soil of nitrophenols, because this practice negatively affected the drinking water supply of Lansing (Colten 1991). And by the 1950s and 1960s industry- and professional associations in western industrial countries were active in promoting practices to prevent soil pollution (Colten 1991; du Mortier et al. 2004). However, court cases and pollution prevention oriented activities have had only limited impact on actual practices. Thus, high levels of contamination are often found at industrial sites where organic compounds were and/or are handled and near waste disposal sites. Also there may be serious soil pollution linked to mishaps in transport and storage. Harm done by soil contamination and limitations to the use of contaminated soils have sparked remediation efforts. Initial government policies directing such efforts during the late 1970s and the 1980s were often aimed at restoration to the original ‘clean’ or pristine state. However as estimated restoration costs for soil pollution mounted, interest in lower cost alternative approaches to soil remediation, such as (monitored) natural attenuation has grown (King and Barker 1999; Renner 2000; Chu et al. 2004; Minsker 2004; Scow and Hicks 2005; McKelvie et al. 2005; Koenigsberg 2006; Mohamed et al. 2006; Schirmer et al. 2006). Natural attenuation has been defined as: naturally occurring processes in soil and groundwater environments that act without human intervention to reduce the mass, toxicity, mobility, and/or volume of contaminants in those media (USEPA 1997). The term natural attenuation is rather often interchanged with terms such as intrinsic (bio)remediation, natural restoration, and self- or passive (bio)remediation. This practice is not followed here as these terms are actually not synonyms (Odencrantz et al. 2003). By now application of (monitored) natural attenuation is also part of a wider policy shift, abandoning the original aim of restoring soils to their original ‘clean’ state in favour of a policy focussed on achieving acceptable or tolerable risk for specified receptors or targets within a ‘reasonable’ time frame (Swartjes 1999; Belluck et al. 2003; Wilson et al. 2004; Nathanail et al. 2005; Tarazona et al. 2005; Provoost et al. 2006; Richardson et al. 2006; Rügner et al. 2006; Atteia and Guillot 2007). The in situ processes mentioned in the USEPA (1997) definition of natural attenuation include biodegradation that tends to be important to natural attenuation of organic contaminants, though other processes including abiotic degradation, trapping, volatilization and sorption can also be significant contributors to thus defined natural attenuation (Hunter et

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al. 1998; Simon and Norris 2006; Vangelas et al. 2006; Chapman et al. 2007; Yoon et al. 2007). It has been pointed out, that there may be objections to the USEPA 1997 definition of natural attenuation. For instance, including volatilization in this definition neglects the potential negative impacts associated with release into the atmosphere, including indoor air. And sorption may be temporary. Against this background it has been proposed to include only destruction and strong sorption by natural processes in the concept of natural attenuation (NRC 2000). This proposal will be followed here. It reinforces the importance of biodegradation in the context of natural attenuation. The founding father of natural attenuation by biodegradation is E.F. Gayle, who in 1952 proposed the microbial infallibility hypothesis, postulating that ‘for any conceivable organic compound there exists a micro-organism that can degrade it under the right conditions’ (Alvarez and Illman 2006). Indeed, a wide variety of micro-organisms has been found in soils and aquifers that can degrade organic contaminants. And factors have been identified that determine whether or not the ‘right conditions’ for biodegradation exist. These include temperature, pollutant and metabolite concentrations, concentrations of electron donors and – acceptors, redox potential, nutrient concentrations, salinity, pH and concentrations of heavy metals (van Hamme et al. 2003; Scott et al. 2005). Monitored natural attenuation is now increasingly advocated, and relied upon, to bring down soil-, and especially groundwater pollution within a ‘reasonable’ time frame. In practice, natural attenuation to achieve current standards, when feasible at all, may be a matter of decades to centuries. When pollution originates in coal tar such natural attenuation may take more than a century, up to millennia (Thornton et al. 2001; Eberhardt and Grathwohl 2002; Hausman and Rifai 2005; Schirmer et al. 2006; Gerhard et al. 2007). In loamy soils, pollution in groundwater by 2,4,6-trinitrotoluene subject to natural attenuation may exceed for millennia current standards (Robertson et al. 2007). Here, against this background, the following questions will be addressed 1. 2. 3. 4.

What is actually known about natural attenuation by biodegradation? With what certainty can natural attenuation by biodegradation be predicted? What does and should ‘tolerable risk’ mean in the context of soil pollution? What do the answers to the previous questions mean for the appropriateness of natural attenuation to bring down high levels of soil pollution to levels of ‘tolerable risk?

2. WHAT IS KNOWN ABOUT BIODEGRADATION IN SOILS? When there is soil pollution with high concentrations of organic compounds, biodegradation of such compounds contributing to natural attenuation can occur. Actual perspectives for biodegradation vary dependent on both the actual organic substance to be degraded and the presence of specific organisms and conditions that are conducive to biodegradation. In practice a number of highly concentrated organic substances that preferentially partition to the particulate fraction of soils have turned out to be to at least a considerable

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extent refractory to biodegradation. And when biodegradation of such compounds occurs, toxic metabolites may be formed. Poor natural attenuation by biodegradation has been found for highly chlorinated compounds such as DDT and polychlorinated biphenyls (PCBs) (Litchfield 2005; Zhang and Bennett 2005). In case of PCBs moreover there is some evidence that metabolites may be toxic to bacteria that can contribute to natural attenuation (Camara et al. 2004). Natural attenuation by biodegradation of soils containing high concentrations of explosives (nitroorganics) has also been found to be poor (Sheremata et al. 2001; Litchfield 2005; Meyers et al. 2007). Both inhibitory effects of pollutants (Meyers et al. 2007) and limited availability to micro-organisms may be involved. Moreover, to the extent that natural attenuation occurs, the presence of N-nitroso intermediates has been noted (Sheremata et al. 2001; Esteve-Nunez et al. 2001). The category of N-nitroso compounds is well known for the carcinogenicity and mutagenicity of many of its members (Esteve-Nunez et al. 2001). The hazardous compound triaminotoluene has been identified as refractory ‘dead-end’ product of anaerobic microbial conversion of nitro-organic explosives (Zhang and Bennett 2005). Polybrominated biphenyls in soils have been found highly resistant to bidegradation (Jacobs et al. 1978). Biotransformation of chlorophenols in soils has been found to mainly lead to the formation of highly persistent chloroanisols and polymeric chlorophenolic compounds (Salkinoja-Salonen et al. 1995). Natural attenuation of hydrocarbons associated with the particulate soil fraction, can occur. Soils with relatively high levels of organic matter seem more conducive to natural attenuation of hydrocarbons than soils with low levels thereof (Scherr et al. 2007). After spills, reductions in concentrations tend to be found for relatively low molecular weight hydrocarbons, but these reductions are probably largely associated with abiotic processes such as volatilization (van Hamme et al. 2003). Branched and cyclic alkanes with 13-25 carbon atoms have turned out to be recalcitrant to biodegradation (Penet et al. 2006). Natural attenuation by biodegradation of major oil spills often levels off at relatively high concentrations of hazardous hydrocarbons (Margesin and Schinner 2001; Mulder et al. 2001; Reddy et al. 2002; van Hamme et al. 2003; Gagni and Cam 2007). Such levelling off may be partly due to inhibition of further biodegradation by metabolites (Chaillan et al. 2006). Susceptibility of polycylic aromatic compounds in soils to biodegradation is severely limited by poor bioavailability and the refractory character of polycyclic aromatics with four or more rings that tends to increase with increasing number of benzene rings (Cassani and Eglinton 1991; Johnsen et al. 2005; Sabate et al. 2006; Gagni and Cam 2007; Soukup et al. 2007). The presence of preferred substrates and the lack of cometabolic inducer substances, metabolite repression and the production of toxic metabolites may contribute to poor biodegradability of polycyclic hydrocarbons (van Hamme et al. 2003). Some of the oxydized metabolites of these hydrocarbons are strong carcinogens (Atlas et al. 1995; Zytner et al. 2006). It would seem that poor availability of pollutants to micro-organisms, conditions that are not conducive to biodegradation, the absence of ‘infallible’ micro-organisms and, in a number of cases, toxicity of pollutants or their metabolites to micro-organisms severely limit natural attenuation of compounds that preferentially partition to the particulate fraction. Natural attenuation to achieve current levels of tolerable risk will take very long time spans, when feasible at all (Zhang and Bennett 2005). During such time spans there will be intolerable risk.

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Better perspectives for natural attenuation by biodegradation have emerged in the context of groundwater remediation. In spite of the microbial infallibility hypothesis, probably not all organic compounds that dissolve in water can be biodegraded. For instance perfluorinated substances are nearly always microbially transformed into extremely persistent perfluorinated acids (Dimitrov et al. 2004), that may contribute to risks on exposure of humans (Betts 2007). But a number of relatively low molecular weight hydrocarbons, oxygenates and chlorinated solvents can in principle be biodegraded (Rittmann 2004). And thus, especially the natural attenuation of low molecular weight aromatic hydrocarbons, chlorinated solvents and the fuel additive methyl-tert-butyl ether in aquifers has attracted much interest (Odencrantz et al. 2003; McGuire et al. 2004; Scow and Hicks 2005). Major advances have been made in understanding natural attenuation of chlorinated solvents: the chloroethenes (Jackson 1998; Lee et al. 1998; Richmond et al. 2001; Bradley 2003; Brungard et al. 2003; McGuire et al. 2004; Smidt and de Vos 2004; Christ et al. 2005; Vangelas et al. 2006; Bhatt et al. 2007; Nijenhuis et al. 2005, 2007). In practice natural attenuation turns out to be very variable. In some cases the concentration of chloroethenes declines strongly, in other cases chloroethenes turn out to be highly conservative (Chapman et al. 2007). Both biodegradation and abiotic degradation (Lee and Batchelor 2002) may occur. At some sites declines in concentration have been found from high levels to levels that meet current standards for tolerable risk (NRC 2000). In oxic groundwater microbial dechlorination of highly chlorinated chloroethenes may occur, but tends to be low. In anoxic groundwater microbial dechlorination of chloroethenes may be more substantial. Efficiency of anaerobic dechlorination appears to decrease however with decreasing chlorine number. More in general, the presence of alternative electron receptors for dechlorinating micro-organisms, low electron donor supply and low effectiveness of electron donors may hamper complete dechlorination. In practice limitations often lead to the accumulation of cis-1,2-dichloroethene and vinylchloride (Skubal et al. 2001; Witt et al. 2002; Dyer 2003). This is a problem because vinylchloride is more toxic than the chlorethenes with three or four chlorine atoms per molecule (Kielhorn et al. 2000; Bronholm et al. 2005). Even under highly reducing conditions the degradation of vinylchloride to non chlorinated compounds proceeds usually at a slow rate and may level off well above tolerable risk levels (Kielhorn et al. 2000; Cupples et al. 2004). Further breakdown of vinylchloride by microbial oxidation is in principle a possibility, but such breakdown is in practice limited by the absence of oxygen in many chloroethene plumes (Bradley 2003). Discharge of chloroethene contaminated groundwater into surface water may be accompanied by considerable biodegradation when low permeability sediments or wetlands have to be passed that are rich in organic matter (Lorah and Olsen 1999; Conant et al. 2004). When there are shallow ponds and small tributary streams in the contaminant plume area upgradient of the river, chloroethenes may be lost from the water phase due to water-air exchange (Chapman et al. 2007). But when this is not the case there may be little attenuation and, when dilution is limited, concentrations in river water may exceed regulatory limits as shown by Ellis and Rivett (2007) for the city of Birmingham and the Tame river in the United Kingdom. Some work has been done on the biodegradation of chlorofluorocarbons, especially CFCs 11 and 12. In this case it has been found in oxic groundwater these CFCs tend to be conservative. However under anaerobic conditions biodegradation may occur, leading to natural attenuation (Jackson 1998; Happell et al. 2003).

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The natural attenuation of relatively low molecular weight aromatic hydrocarbons such as BT(E)X (mixtures of benzene, toluene (ethylbenzene) and xylenes) has attracted much attention as groundwater pollution with these relatively toxic compounds is common. It has been found that biodegradation of BT(E)X is often more substantial than the natural attenuation of chloroethenes (Grandel and Dahmke 2004). Considerable biodegradation of the parent compounds has been found to occur in many field settings, both under aerobic and anaerobic conditions (e.g. Davis et al. 1999; Gieg et al. 1999; Lovley and Anderson 2000; Sueker 2001; Mancini et al. 2003; Richnow et al. 2003; Silva Nunes-Halldorson et al. 2004; Scott et al. 2005; Young and Phelps 2005). Under aerobic conditions also phenols such as catechol may be formed that, in high concentrations, may have an inhibitory effect on natural attenuation (Johnson et al. 2003; Ulrich et al. 2005; Munoz et al. 2007). In case of benzene under sulphate or nitrate reducing and methanogenic conditions benzoate may be generated, which is in turn open to ring cleavage by hydrolysis (Johnson et al. 2003; Ulrich et al. 2005). Furthermore, in case of anaerobic biodegradation benzylsuccinates can be major metabolites (Gieg et al. 1999; Mckelvie et al. 2005; Ulrich et al. 2005). Actual outcomes of natural attenuation turn out to be quite variable, dependent on actual conditions and micro-organisms present. Biodegradation of specific compounds may be strongly dependent on conditions. For instance benzene tends to be poorly degraded under methanogenic conditions and relatively well when conditions are sulfidogenic (Siddique et al. 2007). Though substantial degradation may occur, it has been pointed out that an exclusive focus on the biodegradation of the BTEX may overestimate reduction of toxicity. Substantial residual toxicity has been found after biodegradation of aromatics that is presumably linked to the presence of toxic intermediates and end products such as catechol (Silva NunesHalldorson et al. 2004). Also residual toxicity for humans does not necessarily parallel residual ecotoxicity (Dawson et al. 2007). The natural attenuation of methyl-tert-butyl ether (MTBE) has turned out to be often much less than the natural attenuation of BTEX (Rittmann 2002; Fiorenza and Ritai 2003; Daugherty et al. 2004; Wilson et al. 2004; Scow and Hicks 2005; Martienssen et al. 2006). It has been suggested that, given a similar scale of use, other ether oxygenates such as ETBE, will pose a similar contamination threat to aquifers as MTBE does (Shih et al. 2004).

3. WITH WHAT CERTAINTY CAN FUTURE CONCENTRATIONS BE PREDICTED? Because time spans for natural attenuation may be a matter of decades, centuries or even millennia, one has to make predictions about the future behaviour of contaminants to verify whether the condition of tolerable or acceptable risk to specified receptors or targets will be met within a ‘reasonable’ time frame. Such predictions require modelling of future plume behaviour (Abriola and Chen 1995; Barry et al. 2002; Béranger et al. 2005; Rifai and Rittaler 2005; Alvarez and Illman 2006; Rügner et al. 2006; Gerhard et al. 2007). A considerable amount of work has been done on the modelling of groundwater contamination. Modelling of future behaviour of monitored contaminant plumes tends to extrapolate past behaviour and biodegradation of pollutants into the future. Account should be taken of processes such as solute transport, advection, diffusion, sorption and dispersion.

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Using modelling, it has been determined whether natural attenuation will be capable or incapable of reducing contaminant concentrations to tolerable risk levels (Nobre and Nobre 2004; Schirmer and Butler 2004; Koenigsberg 2006; Atteia and Guillot 2007).

3.1 Determinants of the uncertainty in the prediction of future pollutant concentrations of expanding plumes in groundwater. As one would like to make firm predictions about future concentrations of contaminants, it is important to consider uncertainties regarding future contaminant behaviour. This is all the more important because there is only limited validation of the models used in predicting future concentrations in the field (Abriola 1989; Abriola and Chen 1995; Barry et al. 2002; Abriola 2005). Here we will firstly focus on the case that the contaminant plume is expanding or may do so in the future. Uncertainty linked to the following determinants of future concentrations, as obtained by modelling, will be considered: 1. release characteristics of the pollutant(s) (location, volume, composition and rate), 2. the extent to which sampling adequately characterizes current aquifer contamination, 3. the extent to which sampling is adequate in the characterisation of subsurface, (de)sorption, dispersivity, permeability and, in case of fractured rock, matrix diffusion, 4. degradation kinetics, 5. the presence of other pollutants, 6. future conditions in the aquifer. As to all these determinants uncertainties are in practice inevitable, but uncertainty analysis of remediation strategies applied in the field is rare (NRC 2004).

3.1.1 Release characteristics of the pollutant(s) (location, volume, composition and rate of dissolution) Predictions are sensitive to uncertainty in the source term (Mayer and Endres 2007). In practice, the release characteristics of the pollutant(s) are often highly uncertain (Thornton et al. 2001; Eberhardt and Grathwohl 2002; Grandel and Dahmke 2004; Mayer and Endres 2007). Modelling furthermore, tends to predict near equilibrium concentrations on dissolution of the source, which - in case of dense nonaqueous phase liquids - tends to be much at variance with field observations (Park and Parker 2005). In modelling it is often assumed that the source is immobile after initial infiltration of the soil but actual practice may be at variance with this assumption. Variable water table conditions may change the position of the source and complicate the establishment of the source term and actual releases for hydrocarbons and halogenated solvents that are often the object of natural attenuation (Fretwell et al. 2005; Oostrom et al. 2006). For instance, it has been found that hydrocarbons which due to variation in the water table have ended up in the upper levels of the seasonally unsaturated zone, may act as a persistent source of groundwater

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contamination once the original source of contamination has become depleted or has been removed (Fretwell et al. 2005). Also when sources contain mixtures of nonaqeous phase liquids with large differences in density and relative solubility, remobilisation may occur when mass is lost from the source (Roy et al. 2002).

3.1.2 The extent to which sampling adequately characterizes current aquifer contamination Even when much effort is put into sampling of existing pollution, characterization may be inadequate due to heterogeneity, e.g. with the pollutant bypassing the sampling station by ‘diving’ or following preferential flow paths (Daugherty et al. 2004; Wilson et al. 2004; Wilson et al. 2005; Yenigül et al. 2006). Dense nonaqueous phase liquids, like for instance chlorinated solvents, are in practice often hard to locate due to the physical and chemical complexity of aquifers and due to hysteresis in liquid entrapment and deficiencies in the understanding of migration and sorption processes (Yoon et al. 2007). Thus the characterization of current aquifer contamination may well lead to considerable uncertainty in predicting future concentrations. 3.1.3 The extent to which sampling is adequate in the characterisation of (de)sorption, dispersivity, permeability and matrix diffusion As noted by the National Research Council (NRC 2004), at many sites there is inadequate site characterization to support remediation strategies. Aquifers tend to be systems that are hard to understand in a way that would allow for correct modelling given practical limitations to sampling (Wilson et al. 2005; Koenigsberg 2006). The variation in the soil system in practice prevents complete description leading to uncertainty in hydraulic and other parameters and therefore prediction is inevitably uncertain (Mohamed et al. 2006; van der Keur and Iversen 2006). A case in point relates to geochemical sorption and desorption processes that are an important determinant of aqueous concentrations. In past modelling such processes have been described in terms of linear isotherms. It has been shown that this may cause large errors in predictions (Haws et al. 2006). This has led to the development of more sophisticated models using a much greater number of parameters relevant to (de)sorption. However in practice such parameters can not be independently determined (Haws et al. 20006). Thus prediction retains substantial uncertainty, as far as (de)sorption processes are concerned. Predictions based on modelling are sensitive to assumptions about dispersivity (Thornton et al. 2001). Estimation of dispersivity is difficult (Wilson et al. 2005; Atteia and Guillot 2007). Field scale dispersivities may be several orders of magnitude greater than laboratory scale dispersivities (Mohamed et al. 2006). Plume modelling usually assumes constant dispersivity, but there is evidence that in actual practice transverse dispersivity may be less than constant due to incomplete diffusive mixing (Gaganis et al. 2005). This will lead to underestimates of future pollutant concentrations. Matrix diffusion is an important determinant of solute transport in fractured rock. It has been found that the effective matrix diffusion coefficient in the field tends to be larger than the matrix diffusion coefficient derived from laboratory tests on rock cores (Zhou et al. 2007). Moreover the diffusion coefficient in the field appears to be scale dependent, with different degrees of fractured rock heterogeneity giving rise to variability in the matrix diffusion

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coefficient of two orders of magnitude (Zhou et al. 2007). Again uncertainty due to variability in matrix diffusion is large and increases with time. Hydraulic conductivity generally varies over an order of magnitude but is commonly assumed to be constant in modelling natural attenuation (Odencrantz et al. 2003). The impact of variation in permeability may be quite substantial. In a study of hydrocarbon degradation in porous media El Kadi (2007) has shown that the standard deviation of hydrocarbon concentration may well have the same order of magnitude as the median of results over a long period of time. Uncertainty due to variability in permeability may thus be large and increases with time.

3.4.1 Degradation kinetics Degradation rates (and future plume lengths) have often been estimated on the basis of sampling along the presumable plume centre line. The assumption usually is that the plume is homogeneous. However it may well be that there are in fact a number of relatively discrete plumes of varying flux, reflecting heterogeneity in aquifer and/or sources (Wilson et al. 2004). Often estimated degradation rates are assumed to reflect first or, less often, zero order Monod kinetics (Hunter et al. 1998; Gupta and Seagren 2005; Beyer et al. 2006). In Monod kinetics both substrate depletion and biomass growth and decay are considered. Sampling to establish degradation rates, quite often regards original pollutants, the presence of degradation products, total inorganic carbon or CO2 (Davis et al. 1999; Mayer et al. 2001; Rittmann 2002; Odencrantz et al. 2003; McGuire et al. 2004; Daugherty et al. 2004; Rittmann 2004). Degradation rates derived from concentrations of pollutants following the centreline monitoring approach neglect the variability in solute concentration due to variations in soil emanation rate. The sampling strategy may also be unable to detect biodegradation occurring primarily by fringe processes (Wilson et al. 2004). Moreover reduction of contaminant concentrations or mass in an aquifer may also occur due to (weak) sorption, advection, entrapment and volatilization. In practice there are very little empirical data to make reliable estimates of the importance of the latter processes (Davis et al. 1999; Rittmann 2004; Simon and Norris 2006; Vangelas et al. 2006; Chapman et al. 2007; Yoon et al. 2007). Often a lumped parameter is generated that reflects the aggregate effect of sorption, advection, entrapment, volatilization and biodegradation that then is interchanged with biodegradation (Odencrantz et al. 2003). This will often lead to an overestimate of degradation rate. Stenback et al. (2004) have furthermore shown that degradation rates calculated from measured concentrations of pollutants can vary by a factor 3, depending on choices regarding dispersion coefficients. Moreover the effects of the processes that impact pollutant concentration and mass may in the future be very different from the past. Thus, the reliability of estimates of degradation rates based on estimates of concentration or mass of pollutants usually cannot be determined (Thornton et al. 2001; Odencrantz et al. 2003) and the estimates may be much at variance with the future degradation rates. Establishment of degradation rates on the basis of metabolite concentrations (e.g. Mancini et al. 2003) should often be considered inappropriate as these metabolites are in turn degraded (Sherwood Lollar et al. 2001) or - in some cases - may also have other sources (Bloom et al. 2000; Gray et al. 2002). Estimates based on the presence of total inorganic carbon may lead to estimated degradation rates that are much higher than the true rate. For instance in case of natural

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attenuation of phenolic compounds in a deep sandstone aquifer it has been suggested that, on the basis total inorganic carbon estimates, overestimation of contaminant degradation may be a factor 4-5 (Mayer et al. 2001). An alternative, that has recently gained substantial popularity among researchers, is the monitoring of isotopic carbon compositions (13C/ 12C) of pollutants, as it is presumed that in biodegradation molecular bonds formed by13C will be less easily broken than bonds formed by 12C. In this method use is made of the classical Rayleigh equation, developed for the separation of gases during distillation, but here describing isotopic fractionation: Rt/Ro = f(a-1), in which R is the ratio of the heavy isotope (13C) to the light isotope (12C) at a specified time (0, t), f is the remaining fraction of the isotope at time t, and a is the enrichment factor. On the basis of measured changes in 13C/ 12C ratio (isotopic fractionation) and using the Rayleigh equation combined with a flow and contaminant transport model, the degradation rate of pollutants may be determined. In doing so it is usually assumed that first order Monod kinetics apply. Thus, a first order degradation rate can be obtained that in turn can be used to predict future concentrations of organic pollutants (Bloom et al. 2000; Sherwood Lollar et al. 2001; Sueker et al. 2001; Richnow et al. 2003; Giebler et al. 2004; Peter et al. 2004; Steinbach et al. 2004; McKelvie et al. 2005; Abe and Hunkeler. 2006; Mak et al. 2006). A conceptual problem with using the Rayleigh equation this way, is that the equation is derived for closed systems, whereas groundwater systems are open systems, and that in this equation dispersion is neglected, which may be at variance with the situation in real-life aquifers, leading to underestimates of actual degree of biodegradation and biodegradation rate (Kopinke et al. 2005; Abe and Hunkeler 2006). Also, the Rayleigh equation refers to the degradation of the original pollutant and does not extend to metabolites, which is a mayor limitation when such metabolites are (eco)toxicologically significant. Furthermore, it should also be noted that phase transfer processes, diffusion and interaction with humic substances (Clark and Fritz 1997; Kopinke et al. 2005) may also lead to isotopic fractionation. The latter complication may in turn in practice cause overestimation of biodegradation. Validation efforts of the isotope fractionation approach have shown that actual fractionation due to biodegradation is strongly dependent on the actual micro-organisms, temperature, nutrients and electron acceptors involved in biodegradation (Bloom et al. 2000; Ward et al. 2000; Sueker 2001; Slater et al. 2001; Mancini et al. 2003; Elsner et al. 2005; Nijenhuis et al. 2005). When there is fractionation, a rather large variation in enrichment factors has been noted. In case of compounds such as tetra- and trichloroethene, the variation reported is roughly a factor 5 (Slater et al. 2001; Elsner et al. 2005). It has also appeared that there may be cases of no significant isotopic fractionation during biodegradation. For instance it has been found that, dependent on geochemistry and species involved, there may be no or a substantial isotope fractionation in anaerobic degradation of toluene (Sueker 2001). In a controlled study by Bugna et al. (2004), no significant isotopic fractionation was observed following partial degradation of benzene, toluene, naphthalene, xylenes and decane. Richnow et al. (2003) studied a landfill plume and found an absence of correlation between isotopic fractionation and predictions by the Raleigh equation for 1,2,4-trimethylbenzene and 2-

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ethyltoluene. Also no significant isotopic fractionation was found in studies concerning the biodegradation of polycylic aromatics (Hammer et al. 1998; Griebler et al. 2004). All in all, isotopic fractionation can provide qualitative evidence. Isotopic fractionation may be considered indicative of biodegradation, though a firm conclusion in this respect would only seem possible when it is excluded that other processes, such as interaction with humic substances and diffusion, cause isotopic fractionation. Degradation rates derived from fractionation data would seem uncertain. Uncertainty about degradation rates amplifies with increasing heterogeneity of the aquifer. Aquifers tend to be heterogeneous mixtures of micro environments that vary regarding their conduciveness to biodegradation, leading to substantial biological heterogeneity (Lee et al. 1998; Mohamed et al. 2006). Beyer et al. (2006) estimate that in highly heterogeneous aquifers uncertainty regarding biodegradation may be up to one order of magnitude. This is supported by a study of Lonborg et al. (2006) that focussed on degradation rates for a variety of aromatic compounds originating in a landfill site. They found that degradation rates may differ by about one order of magnitude for Fe- reducing conditions and methanogenic-sulphate reducing conditions. Moreover it may be uncertain whether biodegradation will continue beyond tolerable risk levels. There will be residual fluxes of pollutants that are not subject to further degradation because no biodegrading biomass can be sustained at the residual concentration (Cirpka and Valocchi 2007). This may be especially a problem in case of relatively hazardous substances. As pointed out before, in case of dichloroethene and vinylchloride a threshold has been found below which anaerobic dechlorination could not be sustained (Cupples et al. 2004). For vinylchloride this threshold corresponds, under the conditions studied, with a concentration of 44 micrograms per litre, well above the maximum contaminant level of the US Environmental Protection Agency of 2 micrograms per litre. Also in case of very large spills biodegradation may run out of bioavailable terminal electron acceptors (Huling et al. 2002). Finally, estimated rate constants based on the assumption of first-order Monod kinetics may be uncertain and this may have implications for predicted concentrations. If degradation actually follows not first order Monod but Michaelis-Menten degradation kinetics (that presume apparently constant levels of active microbial biomass involved in degradation), length estimates for contaminant plumes may be less than 40% of their true length (Beyer et al. 2006). And when Monod kinetics is not fist but zero order, biodegradation may be overestimated by a factor of about 4 (Gupta and Seagren 2005).

3.1.5 The presence of other pollutants Complexity of pollution makes predictions about natural attenuation less certain. Uncertainty is in practice compounded by limited analytical completeness, which may lead to missing significant pollutants (Barcelona 2005). Also modelling the biodegradation of mixtures has often not been validated (King and Barker 1999; Reardon et al. 2002; NRC 2004). And remobilisation of source material due to differential mass loss from mixtures is usually neglected (see section 3.1.1). In mixtures present in aquifers pollutants may compete with each other (for e.g. electron acceptors) and as a result thereof plumes of specific organic substances may be or become longer than when only one substance would have been present, as noted by Kampbell and Wilson (1991), Corseuil et al. (1998) Skubal et al. (2001) and Atteia and Guillot (2007). Also toxicity of pollutants or their metabolites to organisms involved in biodegradation may lead to

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longer plumes than expected (Reardon et al. 2002). However there have also been cases in which a mixture led to more rapid degradation or to no difference from the case of a single pollutant (Reardon et al. 2002; Hamed et al. 2003; Zytner et al. 2007). An additional complication arises when aquifers are polluted by more than one source. In this context it has for instance been suggested that MTBE, which is often poorly degraded, may be a cumulative contamination hazard (Reisinger and Reid 2001). Also in case of mining facilities, oil refineries and gas production plants multiple sources of hydrocarbons tend to occur that can give rise to unexpectedly high concentration of hydrocarbons (e.g. Bugna et al. 2004). Similarly multiple sources for halocarbons in aquifers are not unusual (Hunkeler et al. 2004; Ellis and Rivett 2007).

3.1.6. Future conditions in the aquifer Future conditions in the aquifer may be different from the conditions under which biodegradation so far has taken place. In part this follows from existing heterogeneity in aquifers that leads to large differences in conduciveness to biodegradation (see section 3.1.4). But conditions may also change due to pollutants in the aquifer. Obvious examples are the depletion of oxygen linked to the degradation of organic compounds to CO2 that can turn aerobic into anaerobic conditions (Loveley 2001), or running out of electron acceptors in case of very large spills (Huling et al. 2002). When long time spans are involved, changes in climate that lead to changes in aquifer hydrology may also be relevant.

3.2 Overall uncertainty in modelling future pollutant behaviour in expanding plumes Thus, in case of expanding contaminant plumes, modelling of future pollutant behaviour is beset by large uncertainties. These tend to increase when contamination of the aquifer lasts longer. One might expect that such uncertainties can be reduced by inverse modelling (Parker and Islam 2000; Essaid et al. 2003), which seeks to model the development of the current contaminant plume from its release. Reverse modelling is intended to facilitate the choice of conceptual model representation, the determination of model parameters that are the best fit to available data and quantification of the quality of fit (Essaid et al. 2003). For predictive purposes, the choice of conceptual model is very important indeed. However in reverse modelling applied to a field situation (BTEX dissolution and biodegradation at the Bemidji crude oil spill site and the consequences thereof for groundwater contamination) similar fits were found for (all) different conceptual models studied. Long term predictions for the different conceptual models were quite variable. Thus, unfortunately reverse modelling did not lead to reduction of uncertainty about the correctness of model choice (Essaid et al. 2003). It can therefore be expected that current predictions may well be much at variance with actual future concentrations. One might argue that this would not matter to future receptors or targets, when predictions reflect worst case assumptions and thus would err ‘on the safe side’. However from the above it would seem that major uncertainties probably imply that future concentrations and plume length may well be underestimated. This may especially hold in

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case of incomplete source identification and for assumptions that are often made in modelling as to transverse dispersivity, degradation kinetics and the impact of other pollutants

3.3 Predicting pollutant concentrations in steady state or contracting pollutant plumes In some cases of monitored natural attenuation, it has been found that high levels of pollutants are reduced to levels that do not exceed current standards for tolerable risk (NRC 2000). In such cases, supposing that monitoring has been adequate, the main remaining question is whether such standards will also in the future be seen to reflect tolerable risk (see section 4). When in steady state or contracting pollutant plumes current standards are not met, predictions of future pollutant levels are necessary. The uncertainty thereof is smaller than in case of expanding plumes, as a number of factors such as assumptions as to heterogeneity of aquifers and estimates of future dispersivity, permeability and matrix diffusion do not contribute to uncertainty. However major uncertainties as to source characteristics and future degradation remain.

3.4 Conclusions as to uncertainty in predicted concentrations All in all, predictions of future concentration of pollutants in aquifers are uncertain. Uncertainty increases with time and is largest in case of expanding plumes, but in practice may also be significant in case of steady state and contracting plumes. Predictions based on current modelling may well be underestimates. That underestimation of future concentrations is indeed a problem, is supported by the phenomenon that in practice natural attenuation strategies are often failing to deliver their promised outcome (Abriola 2005).

4. TOLERABLE RISK For establishing tolerable risk it is necessary to answer two questions: • •

can actual risk associated with a specified concentration of a pollutant in groundwater or soil be established? how does one determine what risk is tolerable?

4.1 The establishment of risk The establishment of actual risk for a future receptor linked to exposure to a specified concentration of pollutant in soil or groundwater is by no means easy. Three conditions should be met:

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Usually, these conditions can not currently be met. Predictions of future exposure to substances apart from the substances considered in the context of a specified polluted soil, are subject to large uncertainties, and for many chemicals current knowledge about dose-effect relations is at best partial and knowledge about future interactions with other substances is even more limited.

4.2 How does one determine what is tolerable? In the context of natural attenuation the determination of tolerable risk largely refers to the future. If we restrict ourselves to groundwater contamination, where perspectives for natural attenuation are relatively good, time spans in the order of up to thousands of year are possible (see section 1). There are divergent ways to deal with future risks. In economics, it is usual to discount such risks with a certain percentage per year. In doing so, risks in the distant future become vanishingly small. This approach has been objected against by proponents of intergenerational equity or intertemporal fairness (Okrent 1999; Ligner 2003). In practice reasoning on the basis of intergenerational equity often leads to the point of view that current and future receptors should be treated as equals. This in turn would mean that environmental standards reflecting tolerable risk should be equally applied to current and future receptors. However a number of question remains about this operationalization of tolerable risk. These become more pressing when pollution is expected to last for a longer time. As current standards reflecting tolerable risk are often human-centred, one might firstly question whether humans will indeed be affected in the distant future. Perhaps, climate change has led to abandonment of the affected site. Secondly, if there are humans in the future which may be receptors for current pollution, one might wonder whether they would have the same preferences that are relevant to determining what tolerable risk is. Maybe, they would think differently about handling uncertainty. Maybe, they would know more about dose-effect relations or, maybe, it would be easier for them to deal with pollution risks than it is for us. Thirdly, that tolerable risk originating in human activities is currently accepted at all, is strongly related to the fact that benefits accrue from these activities. The benefit of current activities may be argued to be a proper object for discounting and then probably becomes vanishingly small in the distant future. If this is a proper way to look at future risks, then future tolerable risk should also become vanishingly small.

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4.3 Flaws in current practice Notwithstanding the questions outlined in the previous paragraph, one may note that in practice current soil and groundwater quality standards are often used to decide whether or not a predicted future concentration is tolerable. In doing so the supposition is that current and future humans should be treated as equals. Whatever the correctness thereof however, there is also the matter whether we currently know what actual current risks of a specific pollution of the soil are. Current standards cover only part of the chemicals present. For instance, of the volatile organic carbon compounds detected in groundwater samples by the US Geological Survey, 21 were unregulated – with no standards in place (Toccalino and Norman 2006). To the extent that quality standards are in place, they may well have flaws in reflecting current risk. In part this follows from incomplete knowledge about dose-effect relations. The importance thereof can be illustrated by downward revisions of standards of tolerable risk in view of new evidence about toxicity, as in case of trichloroethylene, tetrachloroethylene and MTBE (Simon 2002; Daugherty et al. 2004; Siegel 2006). Estimates of risk may also be flawed due to neglect of interactions between substances in the establishment of current quality standards (Altenburger et al. 2004; Sumpter and Johnson 2005). Even only one pollutant is present, there may be conceptual flaws. For instance in countries like Great Britain, Canada, Germany and Spain soil and groundwater standards are established, while correctly including current background dietary and inhalatory exposure but other countries, like for instance Sweden, Norway and the Netherlands, do neglect such exposure (Tarzona et al. 2005; Provoost et al. 2006). Also current soil and groundwater standards may be dependent on questionable assumptions about exposure routes. In Sweden it is assumed that drinking water is the dominant route of exposure to soil and groundwater pollution, whereas in Norway it is thought that both drinking water and ingestion of soil contribute. In Spain inhalation of soil particles is thought to be relevant to exposure (Tarzona et al. 2005), as confirmed by Nawrot et al. (2006), but in the Netherlands consideration of such inhalation is not included in establishing quality standards. In the USA groundwater standards for volatile organic compounds reflect risk when such water is used as drinking water, but neglect volatilization from groundwater which may lead to airborne risks (Siegel 2006). Dermal uptake is generally neglected, though it has been argued to be relevant for nonionic organic chemicals with a molecular weight less than or equal to 280 (McKone and Howd 1992).

5. THE APPROPRIATE PLACE FOR NATURAL ATTENUATION As pointed out in section 2, the best possibilities for natural attenuation of organic chemicals would seem to exist in aquifers with contaminant plumes. However, especially in cases that the contaminant plume is expanding, predictions whether current standards will be met in the future are highly uncertain and future humans may consider current standards at variance with what they consider to be tolerable risk.

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Thus, and especially when significant contamination is expected to last for tens of years or longer, there seems only very limited scope for using natural attenuation to achieve tolerable risk. This means that, within the context of risk oriented soil pollution policy, often other methods for risk reduction should be contemplated. In these methods a strong focus on source removal would seem necessary whenever the source is still present and removal feasible (Sale and McWhorter 2001; Christ et al. 2005; Brusseau et al. 2007). Such source removal can then be combined with a variety of approaches, with the appropriate approach to be selected on the basis of the actual situation at hand, including the apparent completeness of source removal (Mayer and Endres 2007). In this context biodegradation can still come up for consideration. In case that contaminants are largely associated with the particulate soil fraction, especially for hydrocarbons a variety of in situ and ex situ methods have been developed that lead to much enhanced biodegradation under a variety of climatic conditions and in a variety of soils (Mougin et al. 2002; McCarthy et al. 2004; Zhang and Bennett 2005; Kukyukina et al. 2003; Zhang and Bennett 2005; Perfumo et al. 2007; Scherr et al. 2007; Valentin et al. 2007). Also there seems to be scope for much enhanced biodegradation of nitroaromatics (Zhang and Bennett 2005). In case of aquifer pollution, methods based on source removal coupled with enhanced microbial dechlorination of chloroethenes have been developed (Christ et al. 2005; Zhang and Bennett 2005). Also, many examples of much enhanced biodegradation of BTEX have been documented. (Shieh and Peralta 2005; Zhang and Bennett 2005) and much enhanced biodegradation of MTBE is also feasible (Stocking et al. 2000; Rittmann 2004) In designing methods for enhanced biodegradation, care should be taken to minimize both toxic by-product formation and the mobilization of chemical substances naturally present in soils that may compromise water quality (Barcelona 2005).

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McKelvie, J.R., Linstrom, J.E., Beller, H.R., Richmond, S.A., Sherwood-Lollar, B. 2005. Analysis of anaerobic BTX biodegradation in a subarctic aquifer using isotopes and benzylsuccinates. Journal of Contaminant Hydrology 81: 167-186 McKone, E. and Howd, R.A. 1992. Estimating dermal uptake of nonionic organic chemicals from water and soil: I. Unified fugacity-based models for risk assessments. Risk Analysis 12: 543-557 Meyers, S.K., Deng, S., Basta, N.T., Clarkson, W.W., Wilber, G.G. Long-term explosive contamination in sol: effects on soil microbial community and bioremediation. Soil and Sediment Contamination 16: 61-77 Mohamed, M.M.A., Hatfield, K., Hassan, A.E. 2006. Monte Carlo Evaluation of microbial mediated contaminant reactions in heterogeneous aquifers. Advances in Water Resources 29: 1123-1139 Mougin, C. 2002. Bioremediation and phytoremediation of industrial PAH-polluted soils. Polycyclic Aromatic Compounds 22: 1011-1043 Mulder, H., Breure, A.M., Rulkens, W.H. 2001. Prediction of complete bioremediation periods for PAH soil pollutants in different physical states by mechanistic models. Chemosphere 43: 1085-1094 Munoz, R., Diaz, L.F., Bordel, S., Villaverde, S. 2007. Inhibitory effects of catechol accumulation on benzene biodegradation in Pseudomonas putida F 1 cultures. Chemosphere 68: 244-252 Nathanail, P., McCaffrey, C., Earl, N., Forster, N.D., Gillett, A.G., Ogden, R.. 2005. A Deterministic Method for deriving Site-Specific Human Health Assessment Criteria for Contaminants in Soil. Human and Ecological Risk Assessment 11: 389-410 Nawrot, T., Plusquin, M., Hogervorst, J., Roels, H.A., Celis, H., Thijs, L., Vangronsveld, J., Van Hecke, E., Staessen, J. 2006. Environmental exposure to cadmium and risk of cancer: s prospective population-based study. The Lancet Oncology 7: 119-126 Nobre, R.C.M. and Nobre, M.M.M. 2004. Natural attenuation of chlorinated organics in a shallow sand aquifer. Journal of Hazardous Materials 110: 129-131 NRC (National Research Council) 2000. Natural attenuation for groundwater remediation. Washington DC: National Academy Press NRC (National Research Council) 2004. Contaminants in the Subsurface: Source Zone Assessment and Remediation. Washington DC: National Academy Press Nijenhuis, I., Andert, J., Beck, K., Kastner, M., Diekert, G., Richnow, H.H., Stable isotope fractionation of tetrachloroethene during reductive dechlorination by Sulfurospirillum multivorans and Desulfitobacterium sp. Strain PCE-S and abiotic reactions with cyanocobalamin, Applied Environmental Microbiology 71: 3413-3419 Nijenhuis, I., Nikolausz, M., Köth, A., Fenföldi, T., Weiss, H., Drangmeister, J., Groszmann, J., Kästner, M., Richnow, H.H. 2007. Assessment of the natural attenuation of chlorinated ethenes in an anaerobic contaminated aquifer in the Bitterfeld/ Wolfen area using stable isotope techniques, microcosm studies and molecular markers. Chemosphere 67: 300-311 Odencrantz, J.E., Vogl, R.A., Varljen, M.D. 2003. Natural attenuation rate clarifications: the true picture is in the details. Soil & Sediment Contamination 12: 663-672 Okrent, D. 1999. On Intergenerational Equity and Its Clash with Intragenerational Equity and on the Need for Policies to Guide the Regulation of Disposal of Wastes and Other Activities Posing Very Long Time Risks. Risk Analysis 19: 877-901

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Silva Nunes-Halldorson V., Steiner, R.L., Smith, G.B. 2004. Residual toxicity after biodegradation: interactions among benzene, toluene, and chloroform. Ecotoxicology and Environmental Safety 37: 162-167 Simon, S.E. 2002. EPA’s revised TCE risk assessment likely to affect future remediation standards. Remediation Winter: 1-3 Skubal, K.L., Barcelona, M.J., Adriaens, P. 2001. An assessment of natural transformation of petroleum hydrocarbons and chlorinated solvents at an aquifer plume transect. Journal of Contaminant Hydrology 49: 151-169 Slater, G.F., Sherwood Lollar, B.S., Sleep, B.E., Edwards, E.A. 2001. Variability in carbon isotopic fractionation during biodegradation of chlorinated ethenes: implications for field applications. Environmental Science and Technology 35: 901-907 Smidt, H. and de Vos W.H. 2004. Anaerobic microbial dehalogenation. Annual Review of Microbiology 58: 43-73 Soukup, D.A., Ulery, A.L., Jones, S.. 2007. Distribution of petroleum and aromatic hydrocarbons at a former crude oil and natural gas production facility. Soil and Sediment Contamination 16: 143-158 Steinbach, A., Seifert, R., Annweiler, E., Michaelis, W. 2004. Hydrogen and carbon isotope fractionation during anaerobic biodegradation of aromatic hydrocarbons –a field study. Environmental Science and Technology 38: 609-616 Stenback, G.A., Ong, K.S., Rogers, S.W, Kjartanson, B.H.. 2004. Impact of transverse and longitudinal dispersion on first-order degradation rate constant estimation. Journal of Contaminant Hydrology 73: 3-14 Stocking, A.J., Deeb, R.A., Flores, A.E., Stringfellow, W., Talley, J., Brownell, R., Kavanaugh, M.C. 2000. Bioremediation of MTBE: a review from a practical perspective. Biodegradation 11: 187-201 Sueker, J.K. 2001. Isotope applications in environmental investigations: theory and use in chlorinated solvent and petroleum hydrocarbon studies. Remediation Winter: 5-24 Sumpter, J.P. and Johnson, A.C. 2005. Lessons from endocrine disruption and their application to other issues concerning trace organics in the aquatic environment. Environmental Science and Technology 39: 4321-4332 Swartjes, F.A.. 1999. Risk-Based Assessment of Soil and Groundwater Quality in the Netherlands: Standards and remediation Urgency. Risk Analysis 19:1235-1248 Tarazona, J.V., Fernandez, M.D., Vega, M.M. 2005. Regulation of contaminated soils in Spain. Journal of Soil and Sediments 5,121-124 Thornton, S.F., Lerner, D.N., Banwart, S.A. 2001. Assessing the natural attenuation of organic contaminants in aquifers using plume-scale electron and carbon balances: model development with analysis of uncertainty and parameter sensitivity. Journal of Contaminant Hydrology 53: 199-232 Toccalino, P.L. and Norman, J.E. 2006. Health-based screening levels to evaluate U.S. Geological Survey groundwater quality data. Risk Analysis 26: 1339-1348 Ulrich, A.C., Beller, H.R., Edwards, E.A. 2005. Metabolites detected during biodegradation of 13C-benzene in nitrate reducing and methanogenic enrichment cultures. Environmental Science and Technology 39: 6681-6691 USEPA 1997. Draft interim final OSWER monitored natural attenuation policy. Washington DC, Office of Solid Waste and Emergency Response.

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Valentin, L., Lu-Chau, T.A., Lopez, C., Feijo, G., Moreira, M.T., Lema, J.M. 200t. Biodegradation of dibenzothiophene, fluoranthene, pyrene and chrysene in a soil slurry reactor by the white-rot fungus Bjerkandera sp. BOS55. Process Biochemistry 42: 641648 Vangelas, K., Cahppelle, F.H., Cummings, J., Johnson, P.C., Lovelace Jr., K.A., Nyer, E.K., Norris, B. 2005. Monitored Natural Attenuation Forum: A panel discussion on the use of integrated mass flux and MNA inconsistencies within federal and state agencies. Remediation Winter: 141-152 Vangelas, K.M., Looney, B.B., Early, T.O. , Gilmore, T., Chapelle, F.H., Adams, K.M., Sink, C.H. 2006. Monitored natural attenuation of chlorinated solvents- moving beyond reductive dechlorination. Remediation Summer: 5-23 Van der Keur, P. and Iversen, B.V. 2006. Uncertainty in soil physical data at river basin scale. Hydrology and Earth System Sciences Discussions. 3: 1281-1313 van Hamme, J.D., Singh, A., Ward, O.P. 2003. Recent advances in petroleum microbiology. Microbiology and Molecular Biology Reviews 67: 503-549 Wilson, R.D., Thornton, S.F., Mackay, D.M. 2004. Challenges in monitoring the natural attenuation of spatially variable plumes. Biodegradation 15: 359-369 Wilson, J.F., Newell, C.J., Seaberg, J., Rittmann, B.E., Wiedemeier, T.H., Dickson, W.Z., Haas, P.F.2005. Monitored Natural Attenuation Panel: Use of Modeling to predict MNA and social issues of active remediation versus MNA. Remediation Summer: 121-137 Witt, M.E., Klecka, G.M., Lutz, E.J., Ei, T.A., Grosso, N.R., Chapelle, F.H. 2002. Natural attenuation of chlorinated solvents at Area 6, Dover Air Force Base: groundwater biogeochemistry. Journal of Contaminant Hydrology 57: 61-80 Yenigül, N.B., Hendsbergen, A.T., Effeki, A.M.M., Dekking, F.M. 2006. Detection of contaminant plumes released from landfills. Hydrology and Earth Systems Sciences Discussions 3: 819-857 Yoon, H., Valocchi, A.J., Werth, C. 2007. Effect of soil moisture dynamics on dense nonaqueous phase liquid (DNAPL) spill zone architecture in heterogeneous porous media. Journal of Contaminant Hydrology 90: 159-183 Young, L.Y. and Phelps, C.D. 2005. Metabolic biomarkers for monitoring in situ anaerobic hydrocarbon degradation. Environmental Health Perspectives 113: 62-67 Zhang, C. and Bennett, G.N. 2005. Biodegradation of xenobiotics by anaerobic bacteria. Applied Microbiology and Biotechnology 67: 600-618 Zhou, Q., Liu, H., Molz, F.J., Zhang, Y., Bodvarsson, G.S. 2007. Field scale effective matrix diffusion coefficients for fractured rock: results from literature survey. Journal of Contaminant Hydrology 93: 161-187 Zytner, R.G., Salb, A.C., Stiver, W.H. 2006. Bioremediation of diesel fuel contaminated soil: comparison of individual compounds to complex mixtures. Soil and Sediment Contamination 15: 277-297

In: Environmental Biodegradation Research Focus Editor: B. Y. Wang, pp. 127-158

ISBN: 978-1-60021-904-7 © 2007 Nova Science Publishers, Inc.

Chapter 5

MICROBIAL POLYAROMATIC HYDROCARBON (PAH) BIODEGRADATION IN SUBMERGED SEDIMENT ENVIRONMENTS Yinjie J. Tang1 and James Carothers Jay Keasling Lab, Department of Chemical Engineering, University of California & Virtual Institute of Microbial Stress and Survival, Lawrence Berkeley National Laboratory, Berkeley, CA, USA

ABSTRACT Polycyclic aromatic hydrocarbons (PAHs) in submerged sediments can have potentially carcinogenic effects on human health through the food chain. PAH compounds persist in submerged sediment because of their very low aqueous solubilities, tendencies to adhere to sediment particles and recalcitrance to biodegradation. This chapter covers research topics important for understanding PAH bioremediation in submerged sediments: 1. microbial processes under anaerobic or aerobic sediment conditions; 2. the effects of adding inexpensive environmentally benign substances to stimulate biodegradation or improve PAH bioavailability; 3. methods to characterize microbial, chemical and physical properties in sediment sites; 4. A mathematical model linking the understanding of chemical, physical, and biological activities occurring in the sediment field. Physical capping is frequently used to treat PAH contaminated submerged sediment sites when the site is large volume and has a low contaminant concentration. Recent research suggests that including a physical cap with in situ bioremediation can reduce the volume of cap required to secure a site and provides a potentially economical way to rapidly remediate contaminated sediment sites.

1

Correspondance should be sent to Dr. Yinjie Tang: Jay Keasling Lab, Department of Chemical Engineering, University of California & Virtual Institute of Microbial Stress and Survival, Lawrence Berkeley National Laboratory, Berkeley, CA 94720, USA; Email: [email protected]; Phone: 01-510-495-2628)

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INTRODUCTION Polycyclic aromatic hydrocarbons (PAHs) are a group of about 100 different non-polar compounds with a structure of two or more fused benzene rings that may reside in contaminated sediments, especially marine harbor sediments where creosote was present. Since PAHs have a similar chemical structure to the nitrogenous bases of DNA (flat and aromatic), exposure to PAHs poses a direct threat to fish and other aquatic resources. Intermediate and high-molecular weight PAHs are bio-concentrated and expose humans to a strong carcinogen through the food chain indirectly (Welch, 2001; USEPA Report, 2003). Microbial research in the last 20 years has shown that many bacteria (more than 160 genera) are able to degrade PAHs to derive energy and metabolic building blocks. Water-soluble PAHs can be completely degraded via enzymatic cleavage within several hours (Geiselbrecht et al., 1998). Microorganism populations are numerous in the aquatic environment and able to thrive in wide ranges of environmental conditions, suggesting that in situ bioremediation may be a very effective strategy for reducing large volume hazards in sediments (Lee and Demora, 1999). Molecular ecological approaches have produced genetically-modified species and microbial populations enriched for indigenous PAH degradation capabilities which possess improved PAH biodegradation function in controlled laboratory conditions. However, studies have also demonstrated that these enhanced PAH-degrading microbes are not as effective in experimental field sites due to the presence of other uncharacterized but competing microorganisms and complicated PAH-sediment-water environmental factors (KriegerBrockett et al., 1999; Tang et al., 2005b; Tang et al., 2006). Furthermore, even if laboratory engineered PAH-degrading microbes had been shown to be efficacious in field studies, strict regulations prevent their release into the environment, rendering them useless as agents for bioremediation (USEPA Report, 2003). In the absence of biodegradation, natural physical/chemical PAH attenuation rates in sediment are extremely slow. Because PAHs have low aqueous solubility and tend to adhere strongly to sediment particles, they can persist in sediments for tens to hundreds of years (Harms and Bosma, 1997; Mulder et al. 2001). The fraction of PAHs in an environment that are extractable and bioavailable declines with increasing contact time; a portion of the PAHs becomes unavailable for direct biodegradation in a process called “aging”, where PAH molecules become trapped in the sediment matrix through sorption and diffusion (leading to “aged” PAHs) (Northcott and Jones, 2001a&b) (Figure 1). The attenuation rates of PAHs in sediment are influenced by many physical and chemical environmental conditions including 1) the extent to which the population of microbes in the environment has adapted to metabolize PAHs 2) the bioavailability of the PAHs (i.e. PAH transport/release from the sediment matrix) and 3) the presence or absence of nutrients and electron acceptors needed to support microbial activities. At this point, bioremediation remains slower and less predictable than traditional (and more expensive) physical treatments currently employed by the United States Environmental Protection Agency (USEPA). For example, PAH-contaminated sediments are either mechanically removed or capped in place under a layer of clean soil or sediment by USEPA. Research has shown that the efficacy of in situ biodegradation can be improved, particularly in anaerobic conditions, through the controlled release of electron acceptors or other nutrients (Coates and Anderson, 2000; Tang

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et al. 2005a). However, aged PAHs in sediments remain difficult to degrade and limit the overall rate of biodegradation in the sediment environment.

Figure 1: Concepts of transport and biodegradation models in sediment environments

Here we review results demonstrating that mathematical simulations of contaminant biodegradation and transport in sediment environments can provide useful insights into key processes impacting the efficiency of PAH attenuation. Mathematical simulations are important for facilitating the design of economical strategies for PAH removal and attenuation that combine newer bioremediation technologies with more traditional physical approaches. This chapter discusses phenanthrene (Figure 2) as a model PAH compound and explores five topics:

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Yinjie Tang and James Carothers 1) the properties of PAHs contributing to their sediment reactivity 2) the influence of environmental characteristics on PAH biodegradation rates 3) methods for monitoring intrinsic phenanthrene biodegradation rates in undisturbed sediments 4) approaches for enhancing bioremediation through the addition of nutrients or electron acceptors 5) mathematical models that systematically-describe key processes impacting the efficiency of natural attenuation in sediments.

SECTION 1: PHYSICAL, CHEMICAL AND BIODEGRDATION PROPERTIES OF PAHS 1.1 Chemical properties Polycyclic aromatic hydrocarbons (PAHs) are a group of about 100 different non-polar compounds with a structure of two or more fused benzene rings. PAHs do not burn easily, although they can be broken down slowly upon exposure to sunlight or through reaction with other chemicals such as O2 (Schwarzenbach et al., 1993). Since PAHs have a similar chemical structure (flat and aromatic) to the four DNA nucleobases thymine, cytosine, guanine and adenine, PAHs can intercalate between the “rungs” of DNA molecules, interfering with normal function (Schwarzenbach et al., 1993). Thus, exposure to PAHs poses a threat to fish and other marine resources as well as a potential carcinogenic risk to human health (Welch, 2001). PAHs can be grouped into two categories: “low molecular weight PAHs” and “high molecular weight PAHs”. Low molecular weight PAHs (2~3 benzene rings) are more easily degraded in the environment and therefore exhibit lower toxicity than high molecular weight PAHs. Low molecular weight PAHs tend to be more soluble in water than high molecular weight PAHs. For example, the EPA model compound phenanthrene (USEPA 2003), a tricyclic aromatic hydrocarbon (Figure 2) that may be the most abundant PAH present in sediment, has a solubility in fresh water of 1 mg/L at room temperature, whereas the higher molecular weight perylene (a four ring PAH) has a solubility of only 0.132 mg/L (Futoma et al. 1981). Note, however, that even low molecular weight PAHs have sufficiently poor solubilities in water that they readily separate into an oily phase or adsorb to the sediment.

1.2 Biodegradation properties Microbial research over the last 20 years has shown that more than 160 genera of bacteria, including Pseudomonas, Alcaligenes, Vibrio, Mycobacterium, Comamonas, Rhodococcus, Neptunomonas naphthovorans, cyanobacterium and Cycloclasticus are able to degrade PAHs to derive energy and metabolic building blocks (Berardesco et al., 1998; Nyer et al., 2001). For example, PAH-degrading species of Pseudomonas and Cycloclasticus have been isolated from the sediments of Puget Sound area in Washington State,, which are able to completely degrade soluble phenanthrene and nathrene (Hedlund et al., 1999; Geiselbrecht et

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al., 1998). Biodegradation in fresh or marine sediments by other organisms such as fungi and grazing protozoa has also been documented (Watanabe et al., 2001).

Figure 2: Cleavage of a phenanthrene molecule

Bacteria generally require the mono- or di-oxygenase enzyme system to break PAH rings (Figure 2b) (Kulisch and Vilker, 1991). In the last decade, much new knowledge has been obtained about the anaerobic PAH degraders, and their catabolic pathways, that inhabit sediments (Coates and Anderson, 2000; Wilson and Bouwer, 1997). MacGillivray and Shiaris (1994) found that in coastal sediments bacteria (and not eukaryotic microorganisms) utilizing an anaerobic pathway were the most prominent phenanthrene degraders. Phenanthrene carboxylate isomer is thought to be the initial oxidation product (Ohmoto et al., 1998) under anaerobic conditions; the complete anaerobic metabolic pathway leading from the initial oxidation product is very complicated and has not been fully identified. Despite uncertainty about the specific metabolic pathways involved, recent studies (Rockne et al., 2000 and 2001 and Tang et al., 2005 a, b and Tang et al., 2006) suggest that anaerobic microbial biodegradation has the potential for widespread application as a treatment technology in submerged marine sediments, especially at depths below the surface layer.

1.3 Interaction with sediments PAHs easily adsorb to sediment and seawater colloids because they are extremely hydrophobic, as indicated by their high octanol-water partitioning coefficients (Schwarzenbach et al., 1993). For example, the bimolecular diffusion coefficient of soluble

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phenanthrene in water is 7×10-6 cm2/s. The effective diffusion coefficient for phenanthrene in the pore-space occupied by water in sediment can be as low as 10-9cm2 /s, due in part to the continuous re-partitioning that the PAH undergoes between the water and adsorbed film on the sediment (Wu and Gschwend, 1986). In marine sediments, PAHs are found in a number of different physical states including: 1) solubilized in water 2) patched as solids on sediment particles and 3) sequestered in the pores of sediment aggregates or organic matter (Harms and Bosma, 1997; Mulder et al., 2001). Over periods of years, a large proportion of the PAHs in sediment become trapped by soprtion and diffusion, resulting in “aged” molecules that are unavailable for direct bioremediation (Northcott and Jones, 2001a&b) (Figure 1). In some cases, PAH-degrading microbes can secrete extra-cellular enzymes causing aged PAHs adsorbed in sediments to become more bioavailable (Vetter et al. 1998). In brief, the fate of PAHs in marine sediments depends crucially upon PAH transport processes (sorption, desorption and diffusion), not simply whether the microbial population is adapted for PAH biodegradation.

SECTION 2: FACTORS THAT INFLUENCE MICROBIAL PAH BIODEGRADATION RATES 2.1 Factors affecting microbial biodegradation rates Understanding the influence of environmental conditions on the observed biodegradation rate is essential for the ultimate success of bioremediation strategies. Sediment environments are very complicated and many conditions are reported to influence biodegradation rates (Tang et al., 2005a). Some of these factors alter the rate of microbial uptake and metabolism of the PAHs (i.e. are intrinsic to the cells), and thus depend upon the adaptation of the microbial population to PAH substrates. Other factors affect the rate of contaminant transport to the microorganisms (bioavailability), including the presence or absence of a sorption phase, the mixing intensity (or bioturbation) and the type of the electron acceptors available (Bosma et al., 1997; Lei, et al., 2005). Table 1 summarizes the physical and chemical features of the heterogeneous microbial environment that affect PAH biodegradation rates. The reported biodegradation rates of phenanthrene from different polluted sediment environments are plotted in Figure 3. An empirical model ranking the factors that exert the most influence on biodegradation rates was produced from these published phenanthrene data (Tang et al., 2005a). In order of significance, Tang et al. (2005a) found that phenanthrene biodegradation rates are most affected by 1) adaptation of the bacterial population to PAH degradation 2) the presence of sorption phase 3) the presence of sufficient and suitable electron acceptors and 4) the incubation temperature.

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Table 1. Review of influential factors affecting microbial biodegradation rate (adapted from Tang et al., 2005). Factors Mechanisms and implications References Electron

Aerobic biodegradation is more rapid;

Nyer, 2001; Bitton, 1999;

acceptors

while anaerobic biodegradation is slow

McFarland, 1991; Coates et

and often relies on a consortium of

al., 1996.

bacteria. PAH-degrader

The prior exposure of PAH or other

Geiselbrecht et al., 1996;

population

aromatic co-substrates (toluene) enriches

Ortiz et al., 2003; Hayes et

population of PAH-degraders.

al., 1999; Catallo& Portier, 1992 ;

PAH molecular

2 or 3 ring PAHs are degradable. 4 or 5

Genthner et al., 1997;

structure

ring PAHs are hard to degrade.

Rothermich et al., 2002.

Bioavailability

PAHs bioavailability is limited by low

Pignatello and Xing, 1996;

solubility and strong

Chung and Alexander,

sorption/sequestration in micropores or

1999.

organic matter. Added Nutrients

Temperature

Biodegradation is enhanced only when the

Venosa, et al 1996; Johnson

background nutrients are insufficient.

and Scow, 1999.

Temperature affects enzyme activities and

Nyer, 2001.

thus rates through the metabolic pathway. Pressure

High pressure in deep-sea sediments

Leahy and Colwell, 1990.

reduces biological activity. pH

Most heterotrophic bacteria favor a pH

Chang et al., 2002.

near neutrality.

Salinity

Reported phenanthrene degraders in low salinity estuarine sediments tolerate a wide range of salinities.

Shiaris, 1989a.

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Figure 3: Comparison of phenanthrene biodegradation rate coefficients in polluted sediment. Shiaris, 1989: aerobic slurry experiment, Boston Harbor; Venosa et al., 1996: aerobic macrocosm experiment (blocked), Delaware Bay; Coates et al., 1996 & 1997: anaerobic slurry experiment, San Diego Bay; Rothermich et al., 2002: anaerobic bottle incubation of disturbed sediments with addition of sulfate, Boston Harbor; Tang, 2004: 42 day incubation whole-core injection experiments (micro-aerobic condition), Eagle Harbor sediment; Tang et al., 2005a: 24 day incubation whole-core injection experiments (anaerobic condition), Eagle Harbor sediment.

2.2 Electron acceptors in submerged sediment limiting PAH biodegradation Oxygen is only present in the surface layer of sediments (<1 cm deep) which means that biodegradation in most of the sediment occurs through the use of less energetically-favorable electron acceptors such as sulfate or Fe3+ (Figure 4) (Bakker and Helder, 1993). Consequently, the poor availability of good electron acceptors can limit the ability of microorganisms to metabolize PAHs in situ. Anaerobic PAH biodegradation rates are typically one to two orders of magnitude slower (Chang et al., 2002; Coates et al., 1996 & 1997) than comparable biodegradation rates in aerobic slurries of native marine bacteria (MacGillivray and Shiaris, 1994; Shiaris, 1989 a & b). Phenanthrene biodegradation rates in submerged contaminated sediments are ~two to three orders magnitude slower than rates measured under aerobic slurry conditions (Tang et al., 2006). In particular, natural attenuation of PAHs is diminished by physical capping (a way to isolate the contaminated submerged sediments allowed in some locales) because the availability of electron acceptors (such as oxygen or sulfate) normally received from the water column becomes reduced. If such conditions are present, the addition of electron acceptors to anaerobic sediments, especially

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those involved or amended in capping, could accelerate PAH biodegradation rates and thus serve as a useful strategy for bioremediation (Eriksson et al., 2003; Mcnally et al., 1998; Rothermich et al., 2002).

Figure 4: Electron acceptor zones in marine sediments (after Berner, 1980).

Various electron acceptors, each having different oxidation potentials, in the submerged sediment can support PAH biodegradation. In order of their depth below the sediment-water interface, the predominant electron acceptors available for oxidation of PAHs in anaerobic sediment layers are nitrate, sulfate, and Fe3+ (Berner, 1980). The thermodynamic framework of anaerobic oxidation of PAHs studied by McFarland and Sims (1991) indicates that the electron acceptor nitrate has a high Gibbs free energy for PAH oxidation (-23.73 kcal/eq) and a yield of biomass very close to that for dissolved O2. The stoichiometries and free energies of PAH mineralization reactions using oxygen, nitrate and sulfate as electron acceptors are listed below in order of most-to-least thermodynamically-favorable. C14 H10 +16.5O2→ 14CO2 + 5H2O

∆GΘ (kcal/eq) = -25.28

C14 H10 + 13.2H+ + 13.2NO3- → 14CO2 + 11.6H2O + 6.6N2

∆GΘ (kcal/eq) = -23.73

C14 H10 + 8.3SO42-+ 9H2O → 14HCO3- + 4.13HS- +4.13H2S + 1.6H+

∆GΘ (kcal/eq) = -1.51

Field studies (Hutchins et al., 1998) indicate that bacteria prefer denitrification compared to sulfate reduction when the nitrate supply is not limiting, consistent with the large

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difference in PAH mineralization reaction free energies for nitrate and sulfate. Experiments with enrichments of nitrate-reducing bacterial consortia have shown that their anaerobic PAH biodegradation rates are ~1-2 orders of magnitude higher than the anaerobic PAH biodegradation rates for enrichments of sulfate-reducing bacterial consortia (Rockne et al., 1998 & 2001; Mcnally et al. 1998). However, nitrate is present only in the surface layer of sediments (<15 mm deep) and its concentration is less than 25 µM (Brandes and Devol, 1995) whereas sulfate is very abundant in marine environments (up to 25 mM). The large difference in the concentrations of these electron acceptors causes sulfate reduction to be more widespread in anaerobic sediment biodegradation despite the fact that the nitrate reduction reaction has a lower free energy (McFarland and Sims, 1991). Yet, PAH mineralization by sulfate reduction may not be complete, owing to the comparatively low free energy of the reaction (Rothermich, et al., 2002). Finally, although Fe3+ can support the anaerobic biodegradation of PAHs, Fe3+ supplements should not be used to accelerate PAH mineralization in the environment because iron salts have low aqueous solubility and, most importantly, Fe3+ is likely toxic in concentrations greater than 1 mg/L (Cunningham, 2001).

SECTION 3: MONITORING PAH BIODEGRADATION RATES IN SEDIMENTS 3.1 Measurement of PAH biodegradation To garner the information necessary to design effective bioremediation strategies, PAH biodegradation rates in sediments must be measured. Two types of approaches have been widely used to obtain the rates of microbial PAH uptake and metabolism likely to exist in sediment environments 1) measurement of sediment slurries (measurement-disturbed sediments) and 2) whole core injection methods (undisturbed sediments) (Figure 5).

Figure 5: Schematic of biodegradation rate measurements in sediment slurries or undisturbed sediment (via whole core injection method).

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Owing to the expense of having a statistically significant set of spatially varying sediment samples, PAH biodegradation rates often have been measured in sediment slurries, i.e., slurries are used to homogenize sediments and microbial consortia. Sediment slurries containing a mixture of PAHs or PAH tracers (often radioactive 14C-labeled PAHs) are placed in an aerated shake flask (aerobic conditions) or sealed serum bottles (anaerobic conditions) and incubated. In this experimental setup, the sediments are disturbed and the PAHs thoroughly mixed -- a situation that is very different from what occurs in the field. Problematically, the slurry conditions themselves may stimulate PAH biodegradation rates (Tang et al., 2005a). The whole core injection method, first used by Meyer-Reil (1986) on substrates, and modified by Krieger-Brocket et al. (1997) for use with contaminants, allows pollutant biodegradation rates to be assessed in undisturbed sediment. Here, the sediment is sampled to a particular depth (up to 10cm or more) at fine intervals from a cylindrical core pre-drilled with injection holes (but sealed with a heat-shrinkable plastic sheath). Without mixing or disturbing the sediment, a radiolabeled tracer (14C contaminant) is injected through holes along the sidewall of the core. The whole core, with film-covered injection holes, is then wrapped in parafilm and incubated in an air chamber to measure biodegradation rates under micro-aerobic conditions, or a N2/CO2/H2 gas chamber to measure biodegradation rates under strict anaerobic conditions. The rates at which the microorganisms present in the sediment incorporate radioactive substrate into macromolecules or lose 14CO2 through respiration can be tracked at suitable time intervals (Deming, 1993). In contrast to the contaminant mineralization rates obtained with sediment slurries or agitated sediment-free bacterial cultures, this method is designed to produce rates that more closely approximate those occurring in the submerged sediment at a contaminated site, and for suitable incubation times, produces rates as a function of depth. The whole core injection method should be effective if the incubation periods are short enough (<20 days) to prevent the 14CO2 indicator (in the form of HCO3-) from diffusing throughout the core. The biodegradation and physicochemical loss rates can be followed as an aggregate by tracing the total amount of unreacted PAH with HPLC (Krahn et al. 1993). If 14C-labeled PAHs are used, evolved 14CO2 from biodegradation alone can be trapped with a NaOH solution and quantified by liquid scintillation counting. Although this second method is more sensitive, labeled PAHs must be added at the start of the experiment. Because of this, the measured rates will reflect the biodegradation of fresh PAHs, but not aged PAHs, which are expected to exhibit much slower biodegradation rates. For the purpose of calculating rate coefficients, the kinetics of PAH degradation are conveniently assumed to be first order , although cautionary remarks regarding first-order kinetics are given in Tang et al (Tang et al., 2005a) . If the PAH is labeled with 14C at a single position, the fraction of PAH conversion corresponds to the amount of evolved 14CO2 in an assumed 1:1 stoichiometric ratio. While the authors recognize that some 14C may be incorporated into biomass and intermediate metabolites, assuming 1:1 ratio provides an upper limit to the first-order rate, which is useful for comparing treatment methods. Using endpoint data, the rate coefficients are taken as:

k1 =

− ln(1 − x) Δt

where ∆t is the incubation time and x is the fractional PAH conversion.

(1)

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3.2 Estimation of microbial adaptation to PAH biodegradation in the sediments An initial lag phase in biodegradation rates has often been observed during the measurement of biodegradation rates in the laboratory. For example, Figure 6 shows that the biodegradation of phenanthrene, indicated by 14CO2 recovery, was very low (<3%) during the first 3-6 days at three different measured marine sediment sites in Eagle Harbor, Washington State (Tang, 2004). An acclimation period (from days to months) for PAH microbial biodegradation is often observed when microbes from sediment sites are first exposed to PAHs (Hayes et al., 1999; Tang et al., 2006). Here, acclimatization, likely resulted either 1) from bacteria adapting themselves to growth in the presence of PAH compounds, 2) or from stress and bacterial degeneration during sediment sampling, storage or experimental manipulation.

Figure 6: Depth-averaged 14CO2 recovery (averaged over 0-10cm, n=2) in Eagle Harbor and Blakely Harbor sites (EH: Eagle Harbor, BH: Blakely) (Adapted from Tang 2004).

To estimate the magnitude of fully-developed biodegradation rates (potential maximum biodegradation rates), a nonlinear logistic model can be used to describe acclimation (Bailey and Ollis, 1986). In this model, the biodegradation rate is proportional to the number of active degraders, given as a dimensionless PAH-degrader population factor as a function of time, X(t). The observed biodegradation rate coefficient (k1, day-1) is the product of an overall firstorder biodegradation rate coefficient k10 (day-1) and a PAH-degrader population factor X(t), that is, k1=k10X(t). Under the preceding assumptions, the solutions to the standard differential equations governing this logistic behavior (Bailey and Ollis, 1986) are given below:

PAH Biodegradation in Submerged Sediment

X 0 e β1t dX (t ) = β 1 X (t )(1 − β 2 X (t )) ⇒ X (t ) = dt 1 − β 2 X 0 (1 − e β1t )

139

(2)

where β1 and β2 are dimensionless coefficients of the logistic model, and X0 is the initial biomass. Assuming C and C0 represent the concentration of phenanthrene at any time and time=0 respectively,

dC = −k10 X (t )C dt

(3)

Solving Equation 2 and 3 simultaneously, the results are: k10

− C = (1 − β 2 X 0 + β 2 X 0 e β1t ) β1β 2 C0

(4)

Equation 4 summarizes the effect of an increase in the PAH-degrader population on the rate of PAH disappearance. The overall biodegradation rate constant k10 can be fit to phenanthrene biodegradation rate data using nonlinear function minimization programs (e.g., MATLAB). The maximum biodegradation rate coefficient without the effect of acclimation, k1m, is given by k1m =k10Xmax, where Xmax is the population factor when the PAH-degrader population reaches its maximum.

3.3 Characterization of sediment biogeochemical properties Supporting analyses of the physical and chemical properties of sediments are important for completely understanding differences in PAH degradation rates. Tang et al. (2006) identified a number of different parameters that must be characterized in order to predict the fate of phenanthrene in Puget Sound sediments, including 1) PAH concentrations 2) total organic carbon (TOC) 3) nutrient source (e.g., total nitrogen) 4) electron acceptors (O2, sulfate, nitrate, Fe3+) 5) sediment particle size distribution 6) total bacterial counts 7) sediment porosity and density and 8) sediment temperature and pH. Faster rates of PAH degradation by bacteria are associated with sediments with 1) higher clay content (mediumgrained particles 156-550 μm) 2) greater surface area 3) greater total organic content and 4) higher background PAH contamination (which may induce the population to become adapted to PAH degradation) (Tang et al., 2006). On the other hand, native sediments with large amounts of sand do not support the same abundance of bacteria as sediments with high clay content; sandy sediments show correspondingly lower PAH biodegradation activity (Tang et al., 2006; Schmidt et al., 1998).

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SECTION 4: IN SITU BIOREMEDIATION OF PAHS IN SUBMERGED SEDIMENTS 4.1 In situ bioremediation: a developing technology for cleaning contaminated sites To clean up submerged marine sediments, an effective management strategy requires attention to many details such as regulatory realities, site-specific considerations, costs and stakeholder interests. Highly contaminated terrestrial sites can be rapidly treated by excavation and landfill, incinerating PAHs or by physically desorbing and collecting PAH molecules using solvent flushing or thermal desorption. Note that the latter approach requires a second phase to completely eliminate PAHs and neither method is practical for remediating large scale submerged sediment sites because the costs of such method can be up to $3000 per cubic yard (USEPA report, 2003). In extremely contaminated marine sites, capping, or the placement of a clean sediment barrier, usually dredge spoils, has been used as a temporary solution in locales, such as Puget Sound in Washington State, that permit this method. For large submerged sediment sites with intermediate levels of contamination, monitored natural attenuation or in situ bioremediation is now generally perceived to be effective for treating sediments contaminated with organic pollutants (Lee and Demora, 1999). During bioremediation, a wide variety of contaminants, including PAH compounds, are either destroyed or modified in place by microorganisms, avoiding the release of these harmful compounds into the ecosystem. (Marine Board, 1997).

4.2 Difficulties of in situ bioremediation Data from applications of in situ bioremediation technology have been collected for the cleanup of the 1989 Exxon Valdez oil spill in Alaska (Bragg et al., 1994) and for statistically -designed blocking biostimulation experiments on Delaware Bay beaches (Venosa et al., 1996). Studies show that in situ biodegradation was not very effective in either instance. The main reason for failure seems to be that the natural environment contains diverse uncharacterized organisms and exhibits complicated sediment-water factors not accommodated in the designs for Valdez remediation or the Delaware Bay experiments. Neither genetically-modified species nor populations of microbes enriched with indigenous PAH degraders perform as well in field sites as in the controlled conditions of the laboratory (Watanabe, 2001; Lee et al., 1999: Yu et al, 2005). Other strategies for increasing in situ biodegradation rates in the open environment have made been similarly ineffective. The direct application of either agents that react chemically with PAHs or of surfactants that increase PAH solubility makes only a minor contribution to biodegradation rates. In both cases, the additives are likely lost to the watercourse, or the sediments themselves, before they function as intended (Tang, 2004). It has been challenging to design remediation strategies that are cognizant of economic and time requirements while achieving a cleanup target or sediment quality objective. There are a number of impediments that make it difficult to predict, control and monitor the course of in situ bioremediation. Table 2 summarizes some

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of the current limitations and highlights the need for systematic research, both in the laboratory and the field (Marine Board, 1997). Table 2. In situ bioremediation: effectiveness and limitations. Effectiveness Limitations Research needs (a) less material and (a) Low bioavailability of (a) biodegradation and waste handling PAHs (aged PAH); bioavailability principles and (b) no need for (b) Less biodegradation of high interactions placement sites for molecular weight PAHs; (b) exploration of aerobic and contaminated materials (c) Long treatment duration; anaerobic degradation; (c) favorable public (d) incomplete degradation of (c) Pilot and field demonstration response and PAHs compounds may result in of effectiveness under various acceptability more toxic intermediate biogeochemical sediment (d) lower cost products environments; (d) analysis of cost effectiveness; (e) exploration of combining insitu bioremediation with traditional remediation methods (e.g. capping)

Figure 7: Enhancing capping efficiency by bioremediation.

4.3 Application of in situ bioremediation with capping Capping is the main remediation strategy employed by the USEPA for large volume submerged sites (USEPA report, 2003) (Figure 7). This technology provides physical isolation of the contaminated sediment from the benthic environment, and prevents resuspension and transport of PAHs to the water column and other sites. To date, most in situ capping involves the use of sandy capping materials. Sandy materials have high permeability

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and low organic matter and clay content which limit their ability to reduce contaminant transport into the overlying water column. For these reasons, the cap must be on the order of several feet thick in order to meet performance goals. For example, the caps for superfund sites in Puget Sound, Washington State are 4-20 feet thick. The cost of capping large contaminated sites can be as high as ten million dollars (USEPA report, 2003). Problematically, capping may shut down the supply of oxygen to sediments, resulting in anaerobic environments and PAH compounds that are minimally degraded. As discussed in section 2, oxygen is the preferred electron acceptor for PAH biodegradation. In fact, the fraction of total carbon mineralization by specific electron acceptors in sediments is as follows: O2 (85%)> SO42- (9%) > NO3- (4.5%) > Fe3+ (1%) (Bakker and Helder, 1993). Because oxygen penetration in submerged native sediment is only 3 to 20 mm deep (the average is 10 mm, depending on water depth and hydrological conditions), most buried sediments, including the sediment cap, will be devoid of the oxygen necessary for efficient biodegradation (Brandes and Devol, 1995; Hoehler et al., 1994). To overcome this problem, a new in situ capping and bioremediation technology was suggested by Tang et al., 2005b. Their analyses indicate that the biodegradation process can be supported by the addition of electron acceptors, mainly sulfate and nitrate, to the anaerobic environment. Table 3 summarizes research on stimulating the rate of anaerobic PAH biodegradation via the addition of sulfate and nitrate as electron acceptors. Table 3. Comparison of nitrate and sulfate anaerobic bioremediation studies. Reference Tang et al., 2005a

NO3Controlled release

SO42Controlled release

Rockne et al., 1998

3.5m M

28 mM

Coates et al., 1997

ND

10 mM

PAHs in sediment slurries

Mcnally et al., 1998

60 mg/L

ND

PAHs in bacterial enrichment, aqueous medium

Experiment systems Undisturbed marine sediment With addition of controlled release of nitrate or sulfate materials. PAHs in fluidized bed reactor

Conclusion Biodegradation rates in the sediments increased 2-3 times after continuous supplement of sulfate and nitrate. Anaerobic degradation rate under sulfate reduction was much slower than rates under nitrate reduction. PAH turnover times were 1-2 months under sulfate reducing conditions. PAH turnover times are 12 - 44 hours under nitrate reduction conditions.

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Table 3. Comparison of nitrate and sulfate anaerobic bioremediation studies. Reference Chang et al., 2002

NO321 mM

SO4222 mM

Experiment systems Phenanthrene in soil slurries

Cunningham et al., 2001

85-125 mg/L

70-100 mg/L

BTEX in groundwater at site

Hutchins, 1991

10 mg/L NO3-N

ND

BTEX in batch microcosm

Conclusion Sulfate reducing conditions: k1=0.239 day-1 Nitrate reducing conditions: k1=0.028 day-1 Full destruction of BTEX in 15 months indicates advantage of combined application of nitrate and sulfate. BTEX removed in 76 days; nitrate was consumed in ratios more than stoichiometry; 500 mg/L NO3-N is inhibitory.

ND: no data k1: first order rate constant.

4.4 Enhancement of biodegradation via controlled release of bio-stimulating compounds To support a high rate of anaerobic microbial biodegradation, optimal concentrations of nitrate (1mg-10mg) and sulfate (10 mM-30 mM) should be maintained in local sediments (Boufadel et al., 1999; Tang 2004). In the open environment, adding high concentrations of nutrients in a single instance (immediate release) is ineffective because of “washout”. Even more, high concentrations of nitrate apparently inhibit microbial activities (Tang et al., 2006). On the other hand, by engineering mechanisms to control the release of nitrate and sulfate, ideally in a way that is bacterially-initiated, it should be possible to attain concentrations of additives high enough to support efficient biodegradation for a long period while avoiding toxicity (Figure 8). The presence of slow release forms of oxygen (phosphate-intercalated magnesium peroxide), sulfate (calcium sulfate) or nitrate (cellulose nitrate) has been shown to effectively support anaerobic biodegradation (Tang et al., 2005a; Rothermich et al., 2002; Bach et al., 2005; http://www.environmental-expert.com/index.asp). For sediments lacking nutrients or carbon sources, applying slow release lactate, fertilizers or biopolymers (such as chitosan, a nitrogen and carbon source) can also stimulate in situ biodegradation rates (Xu et al., 2005a and b). Adding controlled release compounds along with the capping materials creates an active bioremediation zone. Because the natural PAH attenuation capacity is enhanced in these zones, the PAHs may be degraded before they are able to diffuse through the cap.

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Consequently, it may be possible to use caps which are much thinner than have typically been required, reducing the overall cost of the cleanup effort by 50~70% (USEPA report, 2003).

Figure 8: Schematic diagram of controlled-release versus immediate release (modified after Ravi Kumar, 2000).

SECTION 5: MODELING PAH BIODEGRADATION IN THE SEDIMENT ENVIRONMENT 5.1 Models describing contaminant fate in submerged sediments Mathematical models are helpful for understanding and predicting how the addition of electron acceptors to in situ bioremediation processes might contribute to the ability to degrade PAH contaminants. Many diagenetic and bioremediation models have been created to 1) evaluate fundamental biodegradation mechanisms in sediments or soils 2) evaluate the fates of contaminants 3) quantify the rate processes and 4) predict the efficaciousness of strategies intended to accelerate bioremediation (Schnoor, 1996; Boudreau, 1997). Shelton and Doherty (1997) described pesticide degradation in soil by coupling Monod biodegradation kinetics with a 2-compartment sorption model in which pollutant bioavailability declines with increasing contact time. Karapanagioti et al. (2001) and Mulder et al. (2001) coupled spherical diffusion with either first-order or Monod biodegradation models to study realistic PAH biodegradation rates in sediments or soils. These models considered how nonlinear contaminant sorption and the physical properties of the system (i.e., density, porosity, particle size and organic matter) affect biodegradation rates. Because additional model parameters, especially those for sorption kinetics, are not easily measured, most studies have focused on a few aspects of the bioremediation process. More complex models are required to describe the biodegradation of organic compounds in soil or sediment media under the constraints of limited oxygen (or other electron acceptors) (Rabouille and Gaillard, 1991; Guha and Jaffe, 1996 and 1999; Brusseau et al., 1999; Zheng et al., 2002).

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5.2 Sorption kinetics model Because bioavailability limits PAH biodegradation rates, a complete model describing their fate in sediments requires correlating microbial biodegradation with contaminant sorption in sediment particulate matter (Brusseau et al., 1999; Karapanagioti et al., 2001; Mulder et al., 2001). Unfortunately, modeling PAH sorption kinetics in sediment is very complicated (Brusseau and Rao, 1989; Harms and Bosma, 1997). Depending on the particular sediment, its organic content and particle size, sediment or soil samples can exhibit either linear or non-linear sorption behavior (isotherms) for contaminants. The Freundlich isotherm model has been useful in this regard because it can describe many types of adsorption sites and contaminant sorption behaviors (Weber and DiGiano, 1996):

S = K PC

n

(5)

Where S is sorption phase concentration (kg kg-1); C is solution-phase contaminant concentration (kg m-3); KP is Freundlich sorption coefficient (m3 kg-1) 1/n; n is Freundlich sorption exponent. In sediments, contaminant sorption and desorption may involve several relevant steps, each with differing rates, such as chemical binding to minerals, diffusion into the pores of particles and aggregates and diffusion within the organic “patches” on the particles (intraorganic diffusion). Such non-equilibrium processes are often described by multi-compartment (also called box or site) kinetics (Brusseau et al., 1989), which requires dividing the sorbent phase into two or more classes of compartments differing in characteristic sorption times. Obviously, model reliability decreases with more uncertain coefficients (Canale and Seo, 1996). The cost associated with obtaining laboratory and field data is very high. It may be best to simplify the model equations and thus reduce the complexity of the model, i.e., the number of independently measured parameters and the number of variables that must be calculated.

5.3 Example of a two dimensional model describing PAH fate in sediment 5.3.1 Model development Here we illustrate an example of a model where the assumption is that the contaminant biodegradation rate is limited by retardation (sorption-retarded diffusion) stemming from an instantaneous equilibrium of Freundlich sorption to sediment particles. Sediment properties and biodegradation parameters are often based on measurements of cylindrical cores; therefore, two dimensional biodegradation and PAH diffusion processes (assuming they obey Fick’s Law) in a cylindrical domain can be described with a dynamic model equation, as outlined by Schnoor (1996):

∂C b ∂ 2C 1 ∂C ∂ 2C = DCε ( + + )−r ∂t ∂r 2 r ∂r ∂Z 2

(6)

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where DC is the contaminant pore fluid effective diffusivity (m2 s-1). The cylindrical geometry of the contaminated site has radius (r) and height (z). The bulk contaminant concentration Cb is computed from the concentrations in the sorption phase and the aqueous phase, using the porosity as a weighting factor:

C

b

= (1 − ε ) ρ s K P C

n

+ εC

(7)

where C is the contaminant concentration in free pore water; ρs is the dried solid phase density; KP is Freundlich sorption coefficient; ε is the porosity; n is the Freundlich constant (for most organic contaminants, n is between 0 and 1). PAH sorption isotherms on some river sediments and coastal sediments are linear, while sorption of PAHs on other soils and shale sites exhibits nonlinearity with a Freundlich constant n as low as 0.19 (Chiou 2002). By coupling equation 6 and an equation for multi-Monod biodegradation with electron acceptor inhibition kinetics for the reaction term r (Nielsen 1994), we extend the general twodimensional transport-reaction equation and obtain:

∂C ∂t

=

ε (1 − ε ) ρ s K P nC

n −1

+ε

[DC (

∂ 2C ∂ 2C 1 ∂C + + ) 2 ∂r ∂Z 2 r ∂r

(8) Ce ] Y C2 Ke + e + Ce KI where Ce is the electron acceptor concentration in pore water (g/L). µmax is the maximum (s-1) bacterial growth rate; X is the biomass concentration (g/L); Y is the yield coefficient; KS and Ke (g/L) are the Monod constants for contaminant and electron acceptors respectively; KI is the electron acceptor inhibition term. From this description, we see that biodegradation will be seriously inhibited by electron acceptors if

−

μ max X

C KS +C

C e2 + C e >> Ke. KI The governing equation for the change of aqueous concentration (Equation 3) can be simplified to a first order reaction model with apparent rate coefficient k1 (the sorption and inhibition retarded coefficient):

μ max X ε dC KI = C dt YKs [( 1 − ε ) ρ S K P + ε ] K I + C e where k 1 =

μ max X ε KI YKs [( 1 − ε ) ρ S K P + ε ] K I + C e

(9) (10)

If the sorption kinetics follow a two-compartment model, then the sorption phase (S) has two compartments (S1 and S2) fed by an aqueous phase PAH (C), and follows first-order

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disappearance kinetics in parallel (Brusseau and Rao, 1989). The governing equations for the contaminant concentrations can be derived as:

∂S1 = k d ,1 [ K PφC − S1 ] ∂t

(11)

∂S 2 = k d , 2 [ K P (1 − φ )C − S 2 ] ∂t

(12)

ε

∂C aq ∂t

∂S1 ∂S ∂ 2 C 1 ∂C ∂ 2 C − ρ S (1 − ε ) 2 + εDC ( 2 + + ) ∂t ∂t r ∂r ∂Z 2 ∂r Ce εμ X C − max Y KS + C C2 K e + e + Ce KI

= − ρ S (1 − ε )

(13)

where ø is in the fraction that is the rapid sorption phase when the sorption compartments are in equilibrium with the aqueous phase; kd,1 is the rate constant (s-1) for the rapidly desorbing fraction; kd,2 is the rate constant for the slowly desorbing fraction (Johnson et al., 2001).

5.3.2. Model dimensionless form Dimensionless forms are applied to simplify the number of independent variables and parameters. The dimensionless variables are Horizontal radius of contaminated site:

r=

r r2

(14)

Depth range of contaminated site:

z=

z z2

(15)

Dimensionless time:

τ =

DC t r12

(16)

Contaminant concentration:

C=

Electron accepter concentration:

C C0 C C e = e0 Ce

Concentration in rapid sorption phase:

S1 =

Concentration in slow sorption phase:

S2 =

S1 φK PC

(17) (18)

0

S2 (1 − φ ) K P C 0

(19) (20)

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where superscript 0 represents the reference concentrations of the contaminants (often contaminant solubility) and the electron acceptors (often their background concentration in the sediment). Table 4 lists the dimensionless parameters describing the PAH biodegradation process in the sediments. Table 4 List of dimensionless parameters in the model equations. Description Geometry and range of contaminated site and bioremediation zone

Sorption

and

diffusion

parameters ( k d ,1 and k d , 2 are similar to the Biot numbers) (Bailey and Ollis, 1986).

Intrinsic parameters

biodegradation

Variables Geometric ratio contaminated site

Dimensionless form of

α0 =

2r2 z2

Spatial factor of active bioremediation zone Partition coefficient

α1 =

r1 r2

Rapid ( k d ,1 ) desorption rate coefficient Slow desorption rate coefficient

K P (1 − ε ) ρ S ( C 0 ) n − 1

w =

ε

2 d ,1 1

k r

k d ,1 =

DC k d , 2 r12

k d ,2 =

DC

Rapid sorption fraction Relative electron acceptor diffusion coefficient Bioremediation stoichiometry

ø

Half saturated Monod constant for contaminant Half saturated Monod constant for electron acceptor

KS =

KS C0

Ke =

Ke C e0

The Damkohler number for the contaminant Electron acceptor inhibition factor

β =

γ =

De DC

C0 γ C e0

Da = KI =

μ max Xr 12 YD C C 0

KI C e0

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149

After transformation, the dimensionless model equations assuming the Freundlich sorption isotherm are:

α 0 α1 ∂ 2 C α 12 Da C ∂ 2 C 1 ∂C ∂C ( ) + − = + n −1 n −1 n −1 2 2 2 ∂τ 1 + nwC r ∂ r 1 + nwC K S +C Ce 1 + nwC ∂z ∂r 2

2

KI

Ce

(21)

+Ce + K e

Dimensionless model equations with two compartment sorption models are: C ∂C ∂ 2 C 1 ∂C ∂2C = α 12 ( 2 + − Da ) + α 02α 12 2 2 ∂τ r ∂r KS + C Ce ∂r ∂z KI

Ce + Ce + K e

− w[φ

∂S 1 ∂τ

+ (1 − φ )

∂S 2 ∂τ

(22)

]

∂S 1

= k d ,1 (C − S 1 ) ∂τ ∂S 2 = k d , 2 (C − S 2 ) ∂τ

(23) (24)

5.3.3 Description of dimensionless model parameters The dimensionless Damkohler number (Da) and the dimensionless partition coefficient describe the key influences on the fate of a contaminant and thus they determine the timeline for bioremediation. Da is proportional to the intrinsic biodegradation rate (

μ max X YC

0

) and

represents the dimensionless bacterial mineralization rate coefficient within the bioremediation zone. The dimensionless partition coefficient, w, represents the ‘severity’ of contaminant sorption, that is, it describes the total sorption capacity of the sediment and the extent to which biodegradation is retarded. w is a function of porosity ε, density ρ and partition coefficient KP. Fig. 9 shows how changes in the dimensionless Damkohler number and partition coefficient affect fresh contaminant biodegradation if the contaminant is not sorbed in the solid phase when time=0. A comparison between linear sorption and nonlinear sorption reveals that nonlinear sorption decreases biodegradation when the contaminant concentration is low, assuming that the contaminant sorption follows the nonlinear Freundlich sorption isotherm (i.e., 0
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in actual sediments or soils (Chung and Alexander, 1999 and 2002; Northcott and Jones, 2001 a, b). Finally, application of those fundamental dynamic models to experimental data is useful for predicting the fate of PAH compounds and estimating the risk of PAHs impacting sediment ecology and human health. However, those models are derived from physical laws and empirical biodegradation kinetics. Complicated environmental factors in the sediment may increase model uncertainties and lead to unmeasured PAH biodegradation and transport processes, including bioturbation, erosion/resuspension, advection, biogeochemical interactions and abiotic transformation reactions (Berner, 1980). For a more advanced analysis, stochastic models should be developed. Such models would use Monte Carlo or bootstrap analysis to generate probability distributions relating microbial attributes and sediment properties with contaminant biodegradation over time and space (Goovaerts et al., 2001).

Fraction of remained contaminant

1

Da=1; w=1

0.9

0.8

Da=10; w=2

0.7

Da=10; w=1

0.6

0.5 0

0.2

0.4

0.6

0.8

1

Time

Figure 9: Effect of dimensionless biodegradation and sorption parameters (Da and w). ( K C =1;

K

I

= 1; β = 1;

γ

= 1; n = 1; α1=1; α0=1;

C

0

=1, and

Ce

0

=1)

Ke

=1;

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Figure 10: Effect of linear (n=1) and nonlinear (0
KC

=1;

Ke

=1;

K

I

= 1; β = 1;

γ

= 0.1; α1=1; α0=1 and

Ce

0

=1)

Figure 11: Effect of only slow sorption on the biodegradation extent under two-compartment sorption model. (Da=10; w=1;

C

0

0

KC 0

=1;

Ke

= S 1 = S 2 =1 and

Ce

=1; 0

K

=1.)

I

= 1;

k d ,1 =1; k d , 2

=0.01; β = 1;

γ

= 0.01; α1=1 ; α0=1;

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SECTION 6: CONCLUSION In summary, PAHs persist in submerged sediments because of their very low aqueous solubilities, tendencies to adhere to sediment particles, and the fact that the nutrients and electron acceptors needed to support microbial degradation are often lacking. However, PAHs are inherently biodegradable in sediment environments if certain conditions are met. Currently, capping is used alone, in some locales, to quickly treat PAH contaminated submerged sediment sites. Recent research has shown that in situ biodegradation can be enhanced through the addition of controlled release electron acceptors and other nutrients. Including these nutrients with a physical cap can reduce the volume of cap required to secure a site, effectively reducing the overall cost. We expect that mathematical simulations of contaminant biodegradation and transport processes in sediments will continue to provide insights into key processes impacting the efficiency of natural PAH attenuation.

NOTATIONS C: solution-phase contaminant concentration, kg m-3 Cb: bulk concentration of contaminant, kg m-3 DC: contaminant pore fluid diffusivity, m2 s-1 De: electron acceptor pore fluid diffusivity, m2 s-1 kd,1 : rate constant for the rapidly desorbing fraction, s-1 kd,2 : rate constant for the slowly desorbing fraction, s-1 KP: Freundlich sorption coefficient, (m3kg-1)1/n KS : contaminant Monod constant, g L-1 Ke : electron acceptor Monod constant, g L-1 n: Freundlich sorption exponent ø: fraction of rapid sorption r1 : radial distance of EA-zone, m; r2 : radial distance of bioremediation zone, m S : sorption phase concentration, kg kg-1 S1 : rapid sorption phase concentration, kg kg-1 S2 : slow sorption phase concentration, kg kg-1 t : time, s X: biomass concentration, kg m-3 Y: yield coefficient, dimensionless z1: the height of EA-zone, m z2: the height of bioremediation zone, m ε: porosity μmax : the maximum growth rate, s-1 ρS: solid phase density, kg m-3 Note: Table 4 shows dimensionless variables.

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ACKNOWLEDGEMENTS The authors thank Professor Barbara Kreiger-Brockett, Professor Jody Deming and Professor Jay Keasling for their advice and support on this manuscript. The authors also gratefully acknowledge the, Washington Sea Grant program, Department of Chemical Engineering and Depatment of Oceanography at the University of Washington, Jane Coffin Childs Memorial Fund, and the US Department of Energy, Office of Biological and Environmental Research, Genomics:GTL Program through contract DE-AC02-05CH11231.

REFERENCES Bach, Q.D., Kim S.J., Choi, S.C., Oh, Y.S. (2005) Enhancing the intrinsic bioremediation of PAH-contaminated anoxic estuarine sediments with biostimulating agents. The Journal of Microbiology 43(4), 319-324. Bailey J.E. & Ollis D.F. (1986) Biochemical Engineering Fundamentals (pp. 388-404). McGraw-Hill Book Company, New York. Bakker, J. F. & Helder, W. (1993) Skagerrak (Northeastern North-Sea) Oxygen Microprofiles and Porewater Chemistry in Sediments, Marine Geology, 111(3-4), 299-321. Berardesco, G., Dyhrman, S., Gallagher, E., & Shiaris, M.P. (1998) Spatial and Temporal Variation of Phenanthrene-Degrading Bacteria in Intertidal Sediments, Applied and Environmental Microbiology, 64(7), 2560-2565. Berner, R. A. (1980) Early Diagenesis: A Theoretical Approach (pp. 81-89). Princeton University Press. Princeton, NJ. Bitton, G. (1999) Wastewater Microbiology (pp. 35-75, 171-206). Wiley-Liss. New York. Bosma, T. N. P., Middeldorp, P. J. M., Schraa G., & Zehnder, A. J. B. (1997) Mass Transfer Limitation of Biotransformation: Quantifying Bioavailability, Environmental Science & Technology, 31, 248-252. Boudreau, B.P. (1997) Diagenetic Models and Their Implementation: Modeling Transport and Reactions in Aquatic Sediments (pp. 59-63, 92-155, 168-189). Springer. New York. Boufadel, M. C., Reeser, P., Suidan, M. T., Wrenn, B. A., Cheng, J., Du, X., Huang, T.H.L., & Venosa, A. D. (1999) Optimal Nitrate Concentration for the Biodegradation of nHeptadecane in a Variably-Saturated Sand Column, Environmental Technology, 20, 191199. Bragg, J. R., Prince, R. C., Harner, E. J., & Atlas, R. M. (1994) Effectiveness of Bioremediation for the Exxon Valdez Oil Spill, Nature, 368, 413-418. Brandes, J. A. & Devol, A. H. (1995) Simultaneous Nitrate and Oxygen Respiration in Coastal Sediments: Evidence for Discrete Diagenesis, Journal of Marine Research, 53, 771-797. Brusseau, M. L., Xie, L. H., & Li, L. (1999) Biodegradation during Contaminant Transport in Porous Media 1: Mathematical Analysis of Controlling Factors, Journal of Contaminant Hydrology, 37, 269-293. Brusseau, M. L. & Rao, P. S.C. (1989) Sorption Nonideality during Organic Contaminant Transport in Porous Media, Critical Reviews in Environmental Control, 19(1), 33-99.

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In: Environmental Biodegradation Research Focus Editor: B. Y. Wang, pp. 159-187

ISBN: 978-1-60021-904-7 © 2007 Nova Science Publishers, Inc.

Chapter 6

MICROBIAL DEGRADATION OF 2-BENZOTHIALZOLE DERIVATIVES: A REVIEW A. Bunescu1, 2, 3 +, P. Besse-Hoggan1, M. Sancelme1, A. Cincilei4, G. Mailhot2 and A.-M. Delort1 1

Laboratoire de Synthèse Et Etude de Systèmes à Intérêt Biologique, UMR 6504 CNRSUniversité Blaise Pascal, 63177 Aubière Cedex, France. E-mail : A-Marie.DELORT@ univ-bpclermont.fr 2 Laboratoire de Photochimie Moléculaire et Macromoléculaire, UMR 6505 CNRSUniversité Blaise Pascal, 63177 Aubière Cedex, France. 3 Laboratory of Organic Chemistry, State University of Moldova, 60, A. Mateevici str, md-2009, Republic of Moldova. 4 Institute of Microbiology, Academy of Sciences, Chisinau, Republic of Moldova. + Present address: Institute of Chemistry, Academy of Sciences, Chisinau, Republic of Moldova.

ABSTRACT This review is focused on one particular family of pollutants, 2-benzothiazole derivatives. This group of xenobiotics containing a benzene ring fused with a thiazole ring is manufactured worldwide. After a short presentation of benzothiazole structures and their industrial applications, the fate of benzothiazoles in the environment is described both in natural waters and in wastewater treatment plants. Then data available on the toxicity of benzothiazoles are reported. The main part of this review is devoted to the microbial degradation of these compounds: i) using activated sludge and mixed cultures, ii) in soils, iii) using pure cultures. In that later case, detailed pathways of biodegradation are described for benzothiazole, 2-hydroxybenzothiazole, 2mercaptobenzothiazole, 2-aminobenzothiazole and methabenzthiazuron. Special attention is made on methodology used to establish these pathways, namely Nuclear Magnetic Resonance (NMR). Finally photodegradation processes are described because molecular mechanisms are often closely related to those of biodegradation processes and lead to

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A. Bunescu, P. Besse-Hoggan, M. Sancelme et al. common products in the environment. To conclude, the possibility of combining these two approaches is discussed.

INTRODUCTION 2-Benzothiazole derivatives constitute a large group of xenobiotics which are manufactured worldwide for a variety of applications (fungicides, herbicides, antialgae agents, colour agents, vulcanisation catalyst, …). They can be recovered in influents of industrial, municipal and domestic wastewater treatment plants, but also in natural environments such as surface waters and soils depending on their use. They present some toxic effects and are thus of concern for the environment and human health. Their chemical structure containing a benzene ring fused with a thiazole ring is rather stable and is quite recalcitrant towards microorganisms and direct photolysis. All these specificities explain why precise metabolic pathways have been described only recently. First studies started with complex media including microbial sludges, soil ecosystems and evolved gradually to isolated and identified mixed or pure cultures of microorganisms. First biodegradation scheme have been explored using radio-labelled molecules. More recently, sophisticated analytical tools were developed including mass spectrometry coupled with different chromatography methods (such as LC-ESI-MS) and Nuclear Magnetic Resonance (in situ one (1D) and two dimensional (2D) experiments), that allowed giving exact structure of metabolites and describing more detailed biodegradation pathways. It was shown that oxidative mechanisms occurred mainly involving mono and dioxygenases and catechol dioxygenases. Oxidation processes involving radical chemistry also take place in photodegradation processes of benzothiazoles; they might contribute, as complementary processes, to the fate of these pollutants in the environment. A new interesting approach could combine photo and biodegradation processes in wastewater treatment plants to improve degradation efficiency. This review is built on two parts: In the first part, the benzothiazole family is presented; it includes the structure and the industrial applications of these compounds, their fate in the environment and data on their toxicity. In the second part are reported biodegradation studies using activated sludge and mixed cultures, in soils and using pure cultures, complementary data about photodegradation pathways are also presented.

PRESENTATION OF THE BENZOTHIAZOLE FAMILY 1. Structure and industrial applications Benzothiazoles are a large family of compounds containing a benzene ring fused with a thiazole ring (Figure 1). They rarely occur as natural products. They form part of the structure of firefly luciferin and are also known as aroma constituents of tea leaves [Vitzthum et al., 1975] and cranberries [Anjou and von Sydow, 1967] or flavour compounds produced by the fungi Aspergillus clavatus and Polyporus frondosus [Seifert and King, 1982; Gallois et al., 1990].

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N H S N

HO

S

S

N

Benzothiazole (BT) Fungicide

H

N

C

S

N COOH

Firefly luciferin

CH3 N CH3

O

Methabenzthiazuron (MBTU) Herbicide

CH3

N

NH2

R

S

HO

N

N O

S

S

Natural flavour produced by fungi

2-Aminobenzothiazole (ABT) Dye

N

N SCH2SCN

SH

S

S

2-Mercaptobenzothiazole (MBT/BTSH) Accelerator of rubber vulcanization

2-(Thiocyanomethylthio)benzothiazole (TCMTB) Fungicide N

N

S

2-Methylthiobenzothiazole (MTBT)

N OH

SCH3 S

2-Hydroxybenzothiazole (OBT or OHBT)

SO3H S

Benzothiazole 2-sulphonate (BTSA)

Figure 1: Chemical structures of some benzothiazoles.

On the other hand, they are worldwide manufactured for a wide variety of applications (Figure 1). They are used as: •

•

• • •

fungicides in lumber and leather production [Reemtsma et al., 1995]: The simplest member of the family, benzothiazole (BT), is a fungicide. 2(Thiocyanomethylthio)benzothiazole (TCMTB) is the active ingredient of the fungicide Busan® [Brownlee et al., 1992] used to protect lumber from fungi attacks [Fiehn et al., 1994]. pesticides and herbicides [Wegler and Eue, 1977; Hartley and Kidd, 1987]: Methabenzthiazuron (MBTU) is used as herbicide in winter corn crops and is an active ingredient of two commercially available formula Tribunil ® and Ormet ®. slimicides in the paper and pulp industry [Meding et al., 1993] in the manufacture of dye : 2-Aminobenzothiazole (ABT) is used in the manufacture of some disperse azo dye [Gaja and Knapp, 1997] chemotherapeutics: Riluzole (2-amino-6-trifluoromethoxybenzothiazole) is marketed by Rhône-Poulenc (Rilutek®) for treatment of amyotrophic lateral sclerosis [Bryson et al., 1996]; 2-(4-aminophenyl)benzothiazole presents antitumour properties [Dubey et al., 2006]. Other derivatives of benzothiazole have been recently developed for their biological activities as inhibitor of topoisomerase

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•

II [Choi et al., 2006] or anti-malaria agents [Hout et al., 2004], some derivatives having similar activities as the reference compounds. Vulcanisation accelerators in rubber production: This application represents the main use of benzothiazole. Benzothiazole derivatives catalyse the formation of sulfide linkages (reticulation) between unsaturated elastomeric polymers in order to obtain a flexible and elastic crosslinked material (Goodyear, Hancock, Parkes processes) (Figure 2).

2-Mercaptobenzothiazole (MBT/BTSH) is the main used rubber accelerator in certain speciality products, and even in the tire production. It can also constitute the starting materials for the synthesis of other derivatives such as 2,2’-(dithiobis)-benzothiazole (MBTS/ BT-S-SBT) [Janin, 1999]. These derivatives have a strong tendency to decompose in the vulcanisation process and can form again MBT [Janin, 1999]. Released in water from manufactured products or from benzothiazole production plants, MBT was found in different environmental compartments with 2-hydroxybenzothiazole (OBT), benzothiazole (BT) and benzothiazole-2-sulphonate (BTSA).

Figure 2: Mechanism of vulcanization [Janin, 1999].

2. Fate of benzothioazoles in the environment Only few data are available about the quantity of benzothiazoles produced each year that can be found in the environment (surface and ground waters) but also in urban wastewater treatment plants. Some data are related to benzothiazoles used for gums and tires (such as MBT). For instance in the 1980’s, the total production of MBT in Europe was estimated to

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38 000 tons [Kloepfer et al., 2005]. In 1985, this production was about 25 000 tons in the USA and the American Environment Protection Agency (www.epa.gov) estimated that about 500 tons were released in nature. The source of benzothiazoles found in the environment or in wastewater treatment plants is diverse. In addition to the production units of these industrial compounds, the wash out of manufactured products (rubber gloves, syringes, tires…) can be a source of indirect contamination. MBT can be released in the environment via stocks of old tires and motorways. It was shown that, after 5 steps of tire wash-out, 50% of the initial content of MBT was extracted [De Wever et al., 2001]. Recently, Kloepfer et al. [2005] showed that domestic wastewaters were also an important source of benzothiazoles, although the exact source remained unknown. Only few studies report quantitative estimation of the amount of benzothiazoles in waters because complex analytical tools are required. Only LC-MS techniques recently adapted by the team of Reemstma [2000] allowed analysing the main benzothiazole derivatives in wastewater treatment plants, in surface waters, in domestic wastewaters and in waters issued from motorway wash-out. We present here the main results of their studies. Figure 3 shows the mean concentrations of MBT, BT and MTBT measured in effluents of two tanneries (I and II) and two pilot water treatment plants (aerobic and anaerobic) of these tanneries [Reemstma et al., 1995]. Actually these three benzothiazole derivatives resulted from the biotransformation of TCMTB used as fungicide for leather treatment. Under anaerobic conditions, MBT was slightly transformed while MTBT decreased and BT increased. Under aerobic conditions, the total concentrations of the three compounds was decreasing, however the authors have shown that this decrease was mainly due to a high adsorption of benzothiazoles on sludge. These results show that these compounds are hardly degradable by biological treatments and can thus be recovered in surface waters. Another study concerns the municipal wastewater treatment plant of Berlin-Ruhleben, where three systems were compared: a system with activated sludge (CAS, Conventional Activated Sludge) and two systems using membrane bioreactors (MBR): MBR1 is functioning with pre-denitrification and MBR2 with post-denitrification [Kloepfer et al., 2004]. The concentrations of ABT, BT, MTBT, BTSA, MBT and OBT in the influent and after treatment in the three systems are reported in Table 1. These results show that the total sum of the benzothiazole derivatives in the influent to be treated was 3.4 μg.L-1 (0.02 μM), so about 300 times less than in the tannery water (see Figure 3). The most concentrated products were BTSA, BT and OBT while MTBT, MBT and ABT concentrations were low. Again the biological treatment systems were rather inefficient on these compounds although great disparities were observed, depending on the studied derivative. In the case of CAS, the total sum of benzothiazoles decreased only slightly with the treatmen. However MBT concentration was decreased by 90%, OBT by 70% (but these two products were minor compounds) while BTSA was increased by 40%. MBR systems were more efficient than CAS; the total sum of benzothiazoles was decreased by 40%. No difference between the two MBR systems was observed.

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Figure 3: Mean concentrations of benzothiazoles measured in tannery effluents before and after treatment in pilot water treatment plants (Number of analysed samples. Tannery I n = 8 ; Tannery II, n = 6 ; total period of observation, 13 weeks) [Reemstma et al., 1995].

Another study conducted in different wastewater treatment plants in Berlin and Beijing [Kloepfer et al., 2005] confirmed the results presented above. The sum of benzothiazoles measured in influent waters varied from 1.9 to 6.5 μg.L-1, again BTSA was the most concentrated derivative (35 to 70% of the total), MBT and ABT concentrations being very low. The monitoring of these stations for a few weeks showed that the decrease of benzothiazole concentration reached only 2 to 5% only. As shown before, BTSA and MTBT were not eliminated but produced. These authors have also measured the fate of these effluents from Berlin in a canal (19 km) linking Berlin to the Tegel lake. These measurements are reported in Figure 4 and showed that benzothiazole concentrations were rather stable in surface waters. Other studies have shown the relative importance of the sources of benzothiazoles entering urban wastewater treatment plants [Kloepfer et al., 2005]. For instance water from road wash-out contained large amounts of benzothiazoles, about 74.5 μg.L-1, which was about 10 times more than in urban wastewater treatment plants. In that case, major benzothiazoles were BTSA (60%), OBT (25-30%) and BT (8-13%), MBT and MTBT were hardly present although MBT is the vulcanization agent for tires. Finally the most unexpected result was the high concentration of benzothiazoles measured in domestic wastewaters, ranging from 50 to 80% of the total amount of benzothiazoles in urban wastewater treatment plants, its origin remained unknown. In conclusion, these studies showed that some benzothiazoles are very recalcitrant and that biological treatments are relatively inefficient for their elimination. Measurements performed in surface waters also show that these compounds are not sensitive to abiotic factors such as solar light. In order to assess the impact of the release of these pollutants in the environment, studies have been carried out to evaluate their toxicity.

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Table 1: Mean concentrations and Standard Deviations (SD) of various benzothiazoles in three wastewater treatment systems and in the influent (mean value of 20 samples collected during 24 hours over a 3 month period). CAS: Conventional Activated Sludge; MBR: Membrane Bioreactors. (Kloepfer et al., 2004). Influent CAS MBR 1 MBR 2 Concentration [ng L-1]

Conc

SD

Conc

SD

Conc

SD

Conc

SD

ABT

27

15

17

5

25

6

15

5

BT

852

197

550

190

286

68

232

85

MTBT

162

61

440

95

284

67

510

180

BTSA

1704

752

2100

484

1517

401

1332

258

MBT

191

71

20

29

21

8

7

7

OBT

501

156

140

79

45

23

75

101

19.8

4.8

17.4

2.9

11.3

2.2

11.3

2.1

58.1

12.7

14.0

2.2

11.9

1.6

11.8

1.4

Sum [nM] -1

DOC [mg L ]

Molar elimination [%] Sum of benzothiazoles

12

43

43

DOC

76

80

80

Figure 4: Concentrations of several benzothiazole derivatives (μg.L-1) along the Blankenfelder graben/Nordgraben canal collecting the effluents of one of the wastewater treatment plant of Berlin. Observation time: 15 h. [Kloepfer et al., 2005].

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3. Toxicity of benzothiazoles Most of the benzothiazole compounds, and particularly MBT, present high toxicity. MBT has a biocide activity towards soil microorganisms at concentrations usually encountered in formulated gums [De Wever, 1995], an antiviral activity [Rada et al., 1979], anti-Candida and anti-fungal properties [Bujdakova et al., 1993]. De Wever et al. [1997a] reported an extensive study of the impact of benzothiazoles on bacteria. They have shown that MBT was the most toxic compound; the thiol group seems responsible of this toxicity [De Wever et al., 1994 a; De Wever, 1995]. MBT is also toxic towards Humans as a powerful allergen; in particular its presence in rubber gloves induces severe dermatitis [Adams and Warshaw, 2006]. Other studies have shown carcinogenic and mutagenic activities in rats and mouses [Gold et al., 1993]. A large study carried out with workers exposed to MBT proved a higher risk of death due to cancer [Whittaker et al., 2004]. Finally a study has been performed in our laboratory to evaluate the ecotoxicity of various benzothiazoles (BT, OBT, MBT, ABT, MBTU, BTSA) and some of their metabolites using the micro-biotest Microtox®. This test is normalized and allows the evaluation of the acute toxicity on a marine bacterium, Vibrio fisheri, which is emitting a natural luminescence. The EC50 value corresponds to the concentration of toxic inhibiting 50% of this luminescence. Table 2 presents the EC50 values obtained for various benzothiazoles and biodegradation products after 30 minutes of exposure to the toxic compound. The more toxic compounds are BT and MBT. It is also clear that the hydroxylated metabolites are always less toxic than the parent compound. Therefore the release of benzothiazoles in the aquatic compartments may be very deleterious for ecosystems and be of concern for the environment and public health. Table 2: Microtox® test values for different benzothiazoles after an exposure time of 30 minutes and corresponding standard deviation (σ) (Malouki, 2004). Benzothiazoles EC50 (mg L-1) σ benzothiazole (BT)

0.364

0.028

2-hydroxybenzothiazole (OBT)

1.801

0.005

2,6-dihydroxybenzothiazole (diOBT)

6.382

0.465

2-mercaptobenzothiazole (MBT)

0.212

0.018

6-hydroxy 2-mercaptobenzothiazole (6OH-MBT)

1.318

0.133

2-aminobenzothiazole (ABT)

6.347

0.713

2-amino-6-hydroxybenzothiazole (6OH-ABT)

16.017

1.442

methabenzthiazuron (MBTU)

22.665

2.385

6-hydroxymethabenzthiazuron (6OH-MBTU)

132.335

12.546

2-benzothiazolylsulfonic acid (BTSA)

294.870

58.260

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BIODEGRADATION OF BENZOTHIAZOLES Although there is an extensive literature on the production and use of benzothiazoles, there are only few reports on their environmental removal from aquatic or terrestrial environments by microflora and even on their biotransformation in laboratory-scale systems [De Wever et al., 2001].

1. Biodegradation using activated sludges and mixed cultures As mentioned previously, the main sources of pollution (or at least the main visible) are issued from factories producing and using benzothiazoles, such as the rubber and tire industries but also tanneries. Therefore the first studies started from the industrial wastewater treatment plants, looking at the efficiency of activated sludges for biotransformation of benzothiazoles in laboratory or pilot-scale systems. The aim of these research works was to improve the process of removal of all the benzothiazole derivatives by controlling the benzothiazole concentration input or by combining physico-chemical and biological treatments for example.

Figure 5: Changes in the concentration of several compounds in the culture filtrate during the growth of a mixed culture in a BTSA-mineral salts medium. Concentration of (S) BTSA ; (z) ammonia ; ({) sulphate ; () organic carbon [Mainprize et al., 1976].

Mainprize et al. [1976] were the first ones to show the complete degradation of BTSA (107 mg.L-1) by activated sludges. In a laboratory-scale activated sludge plant, inoculated with a BTSA-degrading mixed culture and sludge (Figure 5), BTSA disappeared steadily

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from the medium and sulphate and ammonia were produced in a nearly stoichiometric proportions (colorimetric measurements). Moreover the amount of organic carbon in the medium decreased in the same way that the amount of BTSA removed. Thus no significant accumulation of any metabolic intermediates was detected. The BTSA adapted sludge was also able to oxidize BT, OBT and also MBT, but as a slower rate. Other microbial populations not previously exposed to BTSA were also shown to degrade this compound but only after a period of acclimatization. These results were in contradiction with the findings of Chudoba et al. [1977] who studied the biodegradation of several benzothiazole derivatives in a mineral medium with 100 mg.L-1 benzothiazole adapted activated sludge and 100 mg.L-1 of the studied benzothiazole. The authors were monitoring the evolution of Chemical Oxygen Demand (COD) as a measure of biochemical substrate oxidation. Under these conditions, only OBT was biotransformed. The use of mixed microbial cultures, derived from activated sludge of an industrial effluent treatment plant treating effluent from a manufacture of rubber additives, has been studied by Gaja and Knapp [1997] for the biodegradation of benzothiazoles. They observed the biodegradation of BT, OBT and for the first time of ABT (at a concentration of 150 mg.L1 ). With ABT, the release of ammonia and sulphate reached 87 and 100% of the theoretical expected yield. Under these conditions, MBT could not act as a growth substrate for any of the cultures studied but it could be partially oxidized to some extent. However, MBT could be removed as rapidly by heat-killed (sterilized at 121°C for 30 min) as by “live” activated sludge [Gaja and Knapp, 1998]. Nevertheless uninoculated controls show no MBT disappearance. The authors suggested a non-enzymatic process for the removal of MBT but related to the presence of biological materials such as FAD. These first results seemed to prove that the choice of the activated sludges was crucial. Therefore to obtain a very efficient process, De Wever and Verachtert [1994] compared the degradation rate of MBT with MBT-adapted activated sludge (MBT+) and non-MBT history sludge (MBT-) in laboratory-scale fed batch systems. With MBT+ sludge, the first 50 mg.L-1 pulse of MBT was only removed after one month, the second one was degraded in about 20 days and the third one within 6 days. With MBT- sludge, the first pulse of MBT was also removed very slowly (around the same time as with MBT+) but this period was obviously an acclimatization phase, since the following pulses were steadily and completely removed within 6 days (Figure 6). In both cases, the only metabolite identified was BTSA. Its concentration remained relatively stable in time with MBT+ and increased slowly with MBT-. The mixture of both activated sludges MBT+ and MBT- gave the best results with an increase of the degradation rate of MBT immediately after the first pulse (Figure 7). BTSA was formed at a stable concentration and polar compounds accumulated. The addition of a nitrogen source increased even more strongly the rate. However, the higher the concentration of MBT was, the more the biodegradation rate slowed down until the complete inhibition when the MBT concentration reached 100 mg.L-1 (0.6 mM) [De Vos et al, 1993a]. The observed disappearance of MBT was proposed to be due to a chemical oxidation into disulfide at pH < 6, that precipitated in the medium [De Vos et al., 1993b].

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Figure 6: Biodegradation of 50 mg.L-1 of MBT pulses in reactors MBT+ and MBT-. Time courses of MBT (×) and BTSA () concentrations. (∗) Addition of 50 mg.L-1 MBT pulse. All concentrations were determined by UV spectrophotometry.

Figure 7: Biodegradation of 50 mg.L-1 of MBT pulses in reactors with the mixture MBT+/MBT-. Time courses of MBT (×) and BTSA () concentrations. (∗) Addition of 50 mg.L-1 MBT pulse. All concentrations were determined by UV spectrophotometry.

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The problem was getting more difficult when the effluent was composed from a mixture of benzothiazole derivatives. Experiments carried out on laboratory-scale reactors showed that the maximum permissible concentrations for a complete degradation of a BT and MBT effluent were 300 and 10 mg.L-1, respectively. Higher MBT levels led to sludge inactivation and nitrification inhibition [Repkina et al., 1983]. With lower concentrations (at the mg.L-1 scale), Reemstma et al. [1995] observed 10% methylation of MBT into MTBT and a complete degradation of BT. De Wever et al. [1994] showed that in laboratory-scale fed batch and continuous reactors, apparently, MBT and BT mutually stimulated their biological breakdown even at a high concentrations of 150 mg.L-1. On the other hand, MBT clearly inhibited BTSA degradation. Several polar metabolites were detected but not identified during these experiments [De Wever et al., 1994]. Therefore several strategies for the complete removal of benzothiazole derivatives from polluted effluents were tested, in particular using physico-chemical pre-treatments in order to increase the biodegradability of such wastewaters. High energy γ-irradiation [Tölgyessy et al., 1986], adsorption step on activated carbon or on powdered coal [Regula et al., 1983] followed by sonication, filtration and dilution, neutralisation of the wastewaters are some of the processes tested to improve the efficiency of the benzothiazole derivatives removal. None of them were really efficient. The previous addition of a nitrogen or carbon source or phosphates in the wastewaters to stimulate the biological step was also tested. Only the addition of peptones as a nitrogen source and the neutralization of the medium by NH4OH or Ca(OH)2 lead to the improvement of the biodegradation rate of MBT (0.25 mg.L-1.day-1 instead of 0.16 mg.L-1.day-1) [De Wever, 1995].

2. Biodegradation in soils Only the biodegradation of benzothiazole derivatives having herbicide properties have been monitored in soils. Cheng et al. [1978] studied the fate of MBTU in soils using a 14C radiolabelled molecule at specific positions. The degradation into CO2 was very slow and several metabolites were detected. Some of them were identified as demethylation products of the urea function and the others, that were not identified, seemed to be degraded more slowly. The attack on the urea function, in particular on the terminal nitrogen atom, was always preferential to that on the heterocycle. The structure of the lateral chain has therefore a great importance on the biodegradation rate of the compound. The longer the chain was, the slower the biodegradation rate was: -NHCH3-CO-NHCH3 (MBTU) < -NHCH3-CO-NH2 (benzthiazuron) < -NH-CONH2 < -NH2 (ABT). The authors have also tested methylaminobenzothiazole and OBT as potential stable intermediates of MBTU biodegradation. Both compounds were more rapidly biotransformed than MBTU itself. A study under more environmental conditions was carried out by Azam et al. [1988] on a soil from an arid Pakistan region. A positive effect of corn straw amendment on MBTU degradation was observed: with this organic amendment, the percentage of MBTU mineralisation was increased as well as its incorporation to the humic fraction (humic acids and humine). The formation of bound-residues increased with the incubation time and could vary according to the physico-chemical properties of soils, in particular the organic matter content.

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The research works of Printz et al. [1995] on the effect of maize straw amendment in laboratory-scale systems as well as in lysimeter studies confirmed these results. They showed that the radiolabelled straw was rapidly mineralised (around 45% within 45 days) stimulating the microbial activity in soil and increasing the MBTU removal. In the amended soil, the main metabolite of MBTU, the demethylated methabenzthiazuron, was formed in a greater extent. The addition of maize straw lead to the intensification of MBTU degradation but also favoured the formation of bound-residues. Similar effects were obtained in lysimeter studies with slight differences due to temperature variation in real soils or the longer incubation time. In a more general study on the fate of herbicides belonging to the phenylurea pesticide family (such as MBTU), Berger [1999] monitored the transformation of 18 herbicides under various conditions: amended or non-amended soils, sterilized soils inoculated with different pure microbial strains, soil slurries. MBTU was found to be the more persistent herbicide in soils, even when these soils were amended with a nitrogen source or straw. The half-life of MBTU varied from 36 to 59 days depending on the soil tested. Assays with a sterilized soil inoculated with different microbial pure strains gave a percentage of MBTU biodegradation between 7 and 94% depending on the strain tested. It is worth noting a strong substrate specificity for several strains towards MBTU (Rhizopus japonicus and Cunninghamella echinulata), whereas others strains were non specific but also not very efficient. With soil slurry experiments, the transformation rate increased with the lipophilicity of the herbicides studied. In every case, the main biodegradative pathway was the N-demethylation of the urea function but others pathways were present.

3. Biodegradation by pure cultures In order to better understand the metabolic pathways involved in the biodegradation processes of benzothiazoles, pure microbial strains have been isolated. Actually very few strains are described in the literature; they were selected from large screening of microbial banks, from wastewater treatment plants or from mixed cultures. Table 3 summarizes the pure microbial strains described in the literature, which are able to biodegrade or biotransform benzothiazoles. Table 3: Pure microbial strains described in the literature which are able to transform benzothiazoles. Microbial strains Benzothiazoles References Bacteria Pseudomonas sp.

MBT

Drotar et al., 1987

Corynebacterium sp.

MBT

Drotar et al,. 1987

Rhodococcus rhodochrous

BT,

OBT18

MBT

OBT,

ABT,

De Wever et al., 1997b, Besse et al., 2001, Haroune et al., 2001, 2002, 2004

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Rhodococcus erythropolis

BT,

OBT,

BTS1

BTSA

2001, Haroune et al., 2001

Rhodococcus

BT, OBT

Gaja et Knapp 1997, Haroune et al.,

pyridinovorans PA

ABT,

De Wever et al., 1998, Besse et al.,

2002

Fungi Cunninghamella

MBTU

Wallnöefer et al., 1976

Hypocrea pilulifera

MBTU

Goettfert et al., 1978

Aspergillus niger

MBTU

Malouki et al., 2003

echinulata

Biotransformation of MBT MBT is rather recalcitrant and difficult to degrade with pure cultures. Many strains are not able to mineralize MBT but can biotransform it, the major metabolite is usually the 2(methylthio)benzothiazole (MTBT) as shown for instance in the case of Pseudomonas sp. or Corynebacterium sp. [Drotar et al., 1987] (Figure 8). Pseudomonas sp. N N S CH3 SH Corynebacterium sp. S S MBT

MTBT

Figure 8: Biotransformation of MBT by Pseudomonas sp. or Corynebacterium sp.

A very detailed study was carried out in our laboratory using the strain Rhodococcus rhodochrous OBT18 [Haroune et al., 2004]. Resting cells were incubated with MBT (1 mM) and the kinetics of degradation were monitored by HPLC coupled with a UV detector but also by in situ 1H NMR. This NMR approach developed in our group [Delort and Combourieu, 2001; Grivet et al., 2003] gives both qualitative and quantitative data and is performed directly on the incubation media without any previous step of purification. An example of kinetic of degradation followed by in situ 1H NMR is presented in Figure 9. Five metabolites could be detected on these spectra appearing and disappearing at different incubation times. Complementary 2D NMR and MS-spectrometry experiments allowed us to identify the structure of metabolites 1, 2 and 3 (Figure 10). Metabolite 1 is a dihydrodiol, the cis-6,7dihydrodihydroxy-2-mercaptobenzothiazole; metabolite 2 is an open form of the benzene ring and corresponds to a diacid derivative of MBT and metabolite 3 is the 6-hydroxy-2mercaptobenzothiazole (6OH-MBT) resulting from the hydroxylation of MBT at the position 6 of the benzene ring. This work also showed for the first time a partial mineralization of MBT (30%) by a pure culture.

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Metabolite 4

Métabolite 4

Metabolite 5 5 Métabolite 24 h

7h Métabolites Metabolites

Metabolite 3 3 Métabolite 4h

Metabolite 22 Métabolite 3h

Metabolite 1 1 Métabolite 2h MBT

1h

Figure 9: Kinetics of degradation of MBT (1 mM) by resting-cells of Rhodococcus rhodochrous OBT18 monitored by in situ 1H NMR.

Biodegradation of BT and OBT The first detailed investigations of BT metabolism were carried out with Rhodococcus rhodochrous OBT18 and Rhodococcus erythropolis BTS1 by the team of De Wever [1997b, 1998]. Using HPLC equipped with a UV-detector they could detect OBT (2hydroxybenzothiazole) as a first intermediate of the biodegradation. However the lack of performing analytical tools avoided to identify the exact structure of a second intermediate. Using in situ 1H NMR and innovative 2D NMR, namely long range 1H-15N HMBC (Heteronuclear Multiple Bond Correlation) it was possible to identify this second metabolite as the 2,6-dihydroxybenzothiazole (diOBT) (Figure 11A). In that case, as already observed for MBT, the hydroxylation of the benzene ring was specific of the position 6 [Besse et al., 2001].

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H4 H5

C5

H6 HO

C6

C4

C7

C3a C8a

N SH S

H7 OH

Metabolite 2

H4 H5

C5

C4

O HO

N SH S

HO O Metabolite 3

N SH HO

S

6-Hydroxy-2-mercaptobenzothiazole (6OH-MBT)

Figure 10: Structure of metabolites 1, 2 and 3.

Another study was carried out using the same strategy to study the biodegradation of BT by the strain Rhodococcus pyridinovorans PA isolated by the group of Knapp [Haroune et al., 2002]. Again OBT and diOBT were found as metabolic intermediates, but a new product was also identified using 2D NMR 1H-13C HSQC (Heteronuclear Single Quantum Coherence) experiments and mass spectrometry (LC-ESI-MS-MS) as an open-ring product (diacid derivative of OBT) (Figure 11B). After identification of these metabolites, the enzymes involved in the biodegradation process were studied, using enzymatic assays and genetic studies, but also in situ NMR experiments. Indeed this later technique was a very useful tool to investigate the effect of a specific enzyme inhibitor. For instance, in the case of Rhodococcus pyridinovorans strain PA, the identification of the dicarboxylic intermediate derived from BT degradation was consistent with intradiol opening of the benzene ring by a catechol 1,2-dioxygenase. To confirm the type of enzyme, a specific inhibitor, the 3fluorocatechol, was added to the incubation medium. A clear inhibition of the metabolism of BT was then observed by in situ 1H NMR. In addition, a biotransformation of 3fluorocatechol, which is a analogous substrate of the catechol 1,2-dioxygenase substrate, was observed in parallel leading to 3-fluoromuconate by in situ 1H and 19F NMR [Haroune et al.,

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175

2002]. Finally the use of mutants of the Rhodococcus pyridinovorans strain PA showed that the first step of hydroxylation of BT into OBT was catalysed by a different enzyme than the second step of hydroxylation of the benzene ring. H4

B

A

H5

N OH

N OH

O HO

S

HO

C4 C5

S HO

O

Figure 11: Structure of (A) 2,6-dihydroxybenzothiazole (diOBT) and (B) diacid derivative of OBT. 1 0.9 Concentration (mM)

0.8

3

4

N

5

2 6

HO

S1

7

NH2

2-amino-6-hydroxybenzothiazole

6OH-ABT

(A)

0.7 0.6 0.5 0.4 0.3 0.2 0.1 0 0

20 40

60

80 100 120 140 160 180 200 Tim e (h)

(B)

Figure 12: A) Structure of 2-amino-6-hydroxybenzothiazole (6OH-ABT), B) Kinetic of transformation of ABT (0.5 mM) (†) and formation of 6OH-ABT ({) by Rhodococcus rhodochrous OBT18.

Biotransformation of ABT Very few strains accept ABT as substrate. It was shown using 1D and 2D NMR experiments that the strains Rhodococcus rhodochrous OBT18 and Rhodococcus erythropolis BTS1 transformed ABT into its hydroxylated derivative in position 6 of the benzene ring (2amino-6-hydroxybenzothiazole, 6OH-ABT), but this metabolite accumulated in the incubation medium without further degradation (Figure 12) [Haroune et al., 2001]. Conclusion: Metabolism of benzothiazoles by Rhodococcus strains In conclusion, the use of NMR allowed us to describe the biodegradative pathways of four benzothiazoles - BT, OBT, ABT and MBT - by three different Rhodococcus strains. The hydroxylation of the benzene ring at the position 6 is a common step of these pathways. The identification of a diacid compound and 3-fluoromuconate clearly showed that a catechol 1,2-

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dioxygenase was involved in the BT and OBT biodegradation by Rhodococcus pyridinovorans strain PA via the ortho pathway of catechols. The identification of a dihydrodiol in the particular case of MBT shows that the catechol (precursor of the diacid derivative) involved in the metabolic pathway of MBT could be formed from two parallel routes. An hydroxylating dioxygenase is present in Rhodococcus rhodochrous strain OBT18. R = OH, SH N R S

Monooxygenase

Dioxygenase

N

N

R HO

R

S

S

HO

OH Monooxygenase Dehydrogenase

N R S

HO

OH Catechol 1,2-dioxygenase N R

O S

HO

O OH

Mineralization

Figure 13: Metabolic pathways involved in the biodegradation of OBT and MBT by Rhodococcus strains.

Common metabolites were observed in these Rhodococcus strains, suggesting that the mechanisms of benzothiazole biodegradation could be generalized (Figure13). In the first route, the initial benzothiazole is hydroxylated in position 6 of the benzene ring, and then in position 7 by the successive action of two monooxygenases, leading to a catechol. In the second route, the dihydrodiol results from the activity of an hydroxylating dioxygenase, which is then dehydrogenated into a catechol. Finally the aromatic ring is cleaved by a catechol 1,2-dioxygenase to produce a diacid compound.

Biotransformation of MBTU Wallnöfer et al. [1976] studied the biotransformation of the herbicide MTBU by the fungus Cunninghamella echinulata. They showed that methabenzthiazuron was biotransformed into two metabolites M1 and M2. M1 was identified by 1H NMR and mass spectrometry after extraction and purification, as benzthiazuron resulting from a

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demethylation of methabenzthiazuron. M2 corresponded to an hydroxylation of the benzene ring at the position 6. Goettfert et al. [1978] studied the biotransformation of 14C-MBTU by the fungal strain Hypocrea pilulifera isolated from soil. After 7 weeks of incubation, 16% of 14C-MBTU was transformed. Five metabolites were purified by column chromatography and characterized by analytical methods; their structure is reported in Table 4. More recently, Malouki et al. [2003] have shown that the biotransformation of MBTU by the fungus Aspergillus niger stopped after the hydroxylation step of the benzene ring at position 6 (main product) and position 5 (minor product) (Figure 14). Again, in this experiment, the use of 2D NMR 1H-15N HMBC technique allowed the identification of these two metabolites without any ambiguity. It is interesting to stress that again, as in the case of other benzothiazoles with Rhodococcus strains, we observed with these three fungal strains (A. niger, H. pilulifera and C. echinulata) a common step of hydroxylation of the benzene ring at the position 6. It should be noted that none of the three Rhodococcus strains studied previously (R. rhodococcus, R. erythropolis, R. pyridinovorans), which were effective on BT, OBT, ABT and MBT, were able to biotransform MTBU. Table 4: Main metabolites isolated after transformation of methabenzthiazuron (MBT) by Hypocrea pilulifera (Goettfert et al., 1978).

N

R3 N

R1

S

R2

R1

R2

R3

1

H

CH3

CONHCH3 (Methabenzthiazuron)

2

H

CH3

CONHCH2OH

3

H

CH3

CONH2

4

H

H

CONHCH3 (Benzthiazuron)

5

H

CH3

H

6

OH

CH3

CONHCH3

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B)

A)

a)

J H4H5

3 4

J H5H7

J H4H5 4

H4

H7

J H5H7

H5

N U S 3

J N3-H4

4

J N3-H5

A. niger N U HO HO

b)

S 6OH-MBTU (majoritaire) N

3

U S 5OH-MBTU (minoritaire)

3

J N’-CH3

J N-CH3

U=NCH3-CO-NHCH3 Figure 14: A) Metabolites obtained during the incubation of MBTU with A. niger B) 2D 1H-15N HMBC NMR experiments. Cross peaks (spots) show long range correlations between 15N and 1H nuclei. Upper trace spectrum corresponds to correlations between H4 and H5 of the benzene ring with N of the thiazole group. Lower trace spectrum corresponds to correlations of the two methyl groups and the two nitrogen groups in the urea chain.

4. Photodegradation processes The studies of photochemistry of this class of compounds (mainly the pesticides) started with direct photolysis in organic solutions, solutions used for the application of these pesticides. Afterwards the studies have been oriented in aqueous solution with environmental concerns. The understanding of photodegradation processes is complementary to biological ones, and both processes can contribute to the fate of benzothiazoles in the environment. However, at the moment the photodegradation is more in relation with the Advanced Oxidation Processes (AOP’s) of these compounds in water for applications in wastewater treatment plants.

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Phototransformation of TCMTB Brownlee et al. [1992] carried out a study on the fate in the aquatic compartment of the fungicide 2-(thiocyanomethylthio)benzothiazole (TCMTB). Under solar light, TCMTB was rapidly photolysed (life time inferior to 1 day) leading to MBT and traces of benzothiazole (BT). MBT was also photodegraded leading to three different products: BT (28-47%), OBT (4-5%) and traces of a non-identified product. The percentages of formation of these products were dependent on the presence of dissolved organic matter in water. The quantum yields, which are the ratio between the number of transformed molecules divided by the number of absorbed photons by the solution, under solar light, of TCMTB and MBT were estimated to 0.01 and 0.002, respectively. At the end of this study the authors proposed a degradation pathway of TCMTB in the environment (Figure 15) showing the formation of stable products BT and OBT. N SCH2SCN S (TCMTB)

hν H2O

N

N SH +

H S

S (MBT)

(BT)

H2O hν N

N H S (BT)

+

OH S (OBT)

Figure 15: Photolysis of TCMTB (Brownlee et al., 1992).

Phototransformation of MBT Parkanyi and Abdelhamid [1985] studied the photolysis of 2-mercaptobenzothiazole (MBT) and identified the main stable photoproducts. MBT was irradiated in Pyrex reactor in the presence of oxygen. When benzene or toluene were used as solvent, the authors observed the bis-(2-benzothyazolyl)disulfide as the major product of transformation. In solution of acetonitrile, methanol or ethanol, the bis-(2-benzothiazolyl)disulfone was obtained as the intermediate product while benzothiazole-2-sulphonate (BTSA) was the final product of the photodegradation (Figure 16). More recently, Malouki et al. [2004] studied the photolysis at 313 nm of MBT in water. MBT appeared more photoreactive than the majority of benzothiazoles. Actually, its anionic form (pKa = 6.94 ± 0.05) which is 10 times more reactive than the molecular form had a quantum yield of disappearance Φ313nm = 0.02. This phototransformation after the desulfurisation mechanism lead to the formation of BT and OBT. Moreover, the authors compared the phototransformation of MBT in Milli-Q water and in lake water. In this last case, they showed that the disappearance of MBT was 4 times faster showing that chromophoric components of natural water of the lake strongly contributed to the phototransformation of MBT in the aquatic environment.

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hν (pyrex)

N

N

N

hν

S

S

[O]

hν

N

N

SH

S [O], EtOH Et -H2

S

N

hν H2O

SO2

SO2 O2S

S

S hν

N

S

hν

N

SO3H S

S

SO3H S

H2O

N H2SO4 S Figure 16: Direct photolysis of MBT in organic solution [Parkanyi and Abdelhamid, 1985].

This study carried out in pure water and in natural water showed clearly that the photoinduced degradation of benzothiazoles was more efficient that the photolysis. As a consequence, to eliminate efficiently benzothiazoles, different advanced oxidation processes have been tested: Fiehn et al. [1998] studied MBT ozonolyse in pure water or in wastewaters of tanneries. MBT was very reactive with ozone and different products were formed. The three major products observed were BT, benzothiazole-2-sulfite and traces of 2(3H)-benzothiazolone and OBT. These compounds were also degraded with ozone and could be mineralised with a longer treatment. The same results were obtained with the waters from the tanneries but with longer ozonation time, showing that wastewaters containing high amount of MBT should easily be detoxified using ozone. More recently, Valdés and Zaror [2005] improved this process by combining the ozone oxidation with the presence of activated carbon. They obtained a great increase of BT degradation. They showed, with radical trap, that the activation of the reaction was due to oxidation reaction of BT at the surface of activated carbon.

Microbial Degradation of 2-Benzothialzole Derivatives: A Review O

O

N

N C N S

? S

O

O

O H N C N CH3 CH3

N O O

N

S

CH3 H

CH3

HO

N

HO

S

IX O

N C N S

b

H3 C

I

N

O

N N

H

CH3 O

N

O

N

N

H3C

O

H3 C

O

N

N N

H

H3 C

CH3 O

VI

O

H

C N C N C

N CH3

CH3

O VII a

O CH3

N

H2 O H

H V

H3 C CH3

CH3

CH3

X

O N

O

O N C N

O

H IV O

H

CH3

III

a

H

N C N

H

CH3

II

O

N

C H

181

VIII

photolyse de 3g de MBTU (I) dans l'acétone sous oxygène pendant 48 h

b photolyse de 3 g de MBTU dans un mélange eau/acétone (6 :7) sous O2 pendant 83 h

O H 3C

O N

O

H 3C

O

N

O

S

CH3

H

CH 3

C N

H3C

XIII

O

N C

O

H

O

N

O

O

S XIV

CH3 N

H

CH3

H N

O N C

S I

N

O

N

CH 3

c

N C

H

S

CH 3

CH3

CH3 N

H

HSO4

XI H3C O

N

N N

O

O

O CH3 O

IV

H 3C

N

H C N

CH 3

OH O

O

O VII

HO

S O

OH XII

3 g d e MBTU (I) dans water/methanol un mélange eau/méthan (3 : 10) c photolyse c) Photolysis of 3g ofdeMBTU (I) in mixture (3/10)olunder oxygen during 140 h so us courant d'oxygène pendant 140 h

Figure 17: Photooxydation of MBTU at λ > 290 nm a) in acetone, b) in a mixture water/acetone, c) in a mixture water/methanol [Sakriβ et al., 1976].

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Phototransformation of MBTU The methabenzthiazuron (MBTU), herbicide with a high photostability under solar light, has been studied by different authors. Sakriβ et al. [1976] studied the photolysis of MBTU in organic solvent at wavelengths higher than 290 nm. These authors obtained a complex mixture of photoproducts. Among them, 14 have been identified by different analytical methods (1H NMR, MS, UV-visible, IR, …). During the photooxidation, different mechanisms were involved depending on the solvent used: oxidation of the lateral methyl group of the urea chain (II), demethylation (III), opening of the benzene ring (X, XIII, XIV) and different reaction leading to poly-azines (IV, V, VI, VII) (Figure 17). The second study on MBTU, in aqueous solution, by Malouki et al. [2003] showed that its photolysis at λ > 290 nm was a very slow process with no real impact in the environment. Due to this very slow photoreactivity, the same authors studied the degradation of MBTU photoinduced by nitrate and nitrite ions in water [Malouki et al., 2005]. The authors showed that the presence of nitrate ions at a concentration of 0.1 mM increased by a factor of 10 MBTU photodegradation in water. They also showed that the degradation of MBTU photoinduced by nitrite or nitrate ions lead to the formation of numerous products (oxidation of aromatic ring and urea chain, demethylation, nitration, opening of the aromatic ring leading to diacids, dialdehydes, …) showing the non-specificity of the photodegradation process. Mineralization was achieved after prolonged exposure. Phototransformation of benzothiazoles : Photo-Fenton process Andreozzi et al. [2000, 2001] also studied the degradation of different benzothiazoles in aqueous solution by a photo-Fenton process (H2O2, iron and light) in a closed reactor. First, they developed a kinetic model, which can predict the disappearance of BT taking into account the influence of pH, concentrations of H2O2 and Fe(III) or the ionic strength of the medium. The model gave results in good correlation with the experimental results except for pH higher than 3.0. The kinetics of aqua-complexes formation and the precipitation of iron hydroxides at such pH are very difficult to take into account [Andreozzi et al., 2000]. This study has been extended at different benzothiazoles like MBT and OBT. This model leads to a good estimation of the kinetic rate constants of the attack of OH radical on the different benzothiazoles. With the modulation of the concentrations of H2O2 and Fe(III) and of the pH, a total disappearance of benzothiazoles was obtained. These results showed that this process is very interesting for the treatment of water polluted by benzothiazoles. To illustrate these results, a very recent and original study has been carried out on the photo-Fenton treatment of leaching water from stock of old tires [Sarasa et al., 2006]. The main organic products identified in this leaching water are: phenol derivatives, phthalates, fatty acids, hydrocarbons and of course benzothiazole derivatives. The photo-Fenton treatment leads to the complete elimination of benzothiazole derivatives. For the complete treatment of this leaching water, the authors showed that the better elimination of organic carbon was obtained after the photo-Fenton treatment followed by a coagulation-flocculation process involving also ferric salts. These few studies on the phototransformation of benzothiazole derivatives show their slow photolysis in aqueous solution. However, the studies involving photoinduced processes (Fenton, Nitrates, …) show a better efficiency for the elimination of benzothiazoles in water.

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CONCLUSION The main conclusion of this paper is that benzothiazoles are not very well bio-or photodegraded by classical methods used in wastewater treatment plants. This class of compounds is recalcitrant to elimination processes due to natural bio- and photodegradation. However, with the goal of increasing the efficiency of benzothiazole elimination in water, photochemists and microbiologists carried out an original study on the sequential bio- and phototransformation of MBTU in water [Malouki et al., 2003]. As mentioned before MBTU was very slowly phototransformed at wavelengths higher than 290 nm but it was successfully oxidized into 6-hydroxymethabenzthiazuron by Aspergillus niger. However, this first metabolite was no further metabolised by Aspergillus niger but was efficiently photooxidized with ring cleavage of the aromatic ring and photodimerized under irradiation at λ > 290 nm. This interesting work shows the complementarity of both approaches, photo- and biodegradation, to study the fate or the elimination of benzothiazoles in water. Combining processes of photo- and biodegradation, successively or simultaneously, could be very promising for recalcitrant pollutants. In addition this strategy could be very interesting from an economical point of view as biotechnological processes and the use of solar light are very inexpensive compared to advanced oxidation processes. This research activity is an important perspective for the future years.

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Malouki, M. (2004). Photo et/ou biotransformation de l’ioxynil et des dérivés benzothiazoliques. Thèse de l’Université Mentouri de Constantine, (République Algérienne Démocratique et Populaire). Malouki, M.A., Richard, C. & Zertal, A. (2004). Photolysis of 2-mercaptobenzothiazole in aqueous medium: Laboratory and field experiments. J. Photochem. Photobiol. A: Chem., 167, 121-126. Malouki, M.A., Lavédrine, B. & Richard, C. (2005). Phototransformation of methabenzthiazuron in the presence of nitrate and nitrite ions. Chemosphere, 60, 15231529. Meding, B., Toren, K., Karlberg, A.T., Hagberg, S. & Wass, K. (1993). Evaluation of skin symptoms among workers at a Swedish paper mill. Am. J. Ind. Med., 23, 721-728. Párkányi, C. & Abdelhamid, A.O. (1985). Photodegradation of pesticide : photolysis of 2mercaptobenzothiazole. Heterocycle, 23, 2917-2926. Printz, H., Burauel, P. & Führ, F. (1995). Effect of organic amendment on degradation and formation of bound residues of methabenzthiazuron in soil under constant climatic conditions. J. Environ. Sci. Health, B30, 435-456. Rada, B., Holbova, E., Mikulasek, S., Sidova, E. & Gvozdjakova, A. (1979). Antiviral activity of benzothiazole and benzothiazolinethione derivatives in cell cultures. Acta Virol., 23, 203-209. Reemtsma, T., Fiehn, O., Kalnowski, G. & Jekel, M. (1995). Microbial transformations and biological effects of fungicide-derived benzothiazoles determined in industrial wastewater. Environ. Sci. Technol., 29, 478-485. Reemstma, T. (2000). Determination of 2-substituted benzothiazoles of industrial use from water by liquid chromatography/electrospray ionization tandem mass spectrometry. Rapid Commun. Mass Spectrom., 14, 1612-1618. Regula, S., Ondris, L. & Kacani, S., (1983). Physicochemical pretreatment of wastewaters from the production of benzothiazole derivatives. Czech. CS 208, 626 December 1st (in Czech). Cited in Chem. Abstr. CA 101 : 136470q. Repkina, V.I., Dokudovskay,a S.A., Umrikhina, R.A. & Samokhina V.A. (1983). Maximum permissible concentrations of benzothiazole and 2-mercaptobenzothiazole during biochemical treatment of wastewaters. Khim. Prom-st., 10, 598-599. Sakriβ, W., Gäb, S. & Korte, F. (1976). Photooxidationsreaktionen von methabenzthiazuron in Lösung. Chemosphere, 5, 339-348. Sarasa, J., Llabrés, T., Ormad, P., Mosteo, R. & Ovelleiro, J.L. (2006). Characterization and photo-Fenton treatment of used tires leachate. J. Hazard. Mater., 136, 874-881. Seifert, R.M. & King, D.A. Jr. (1982). Identification of some volatile constituents of Aspergillus clavatus. J. Agric. Food Chem., 30, 786–790. Tölgyessy, P., Kollar, M., Vanco, D. & Piatrik, M. (1986). The radiation treatment of waste water solutions containing 2-mercaptobenzothiole and N-oxydiethylene-2-benzothiazole sulfenamide. J. Radioanal. Nucl. Chem., 107, 315-320. Valdés, H. & Zaror, C.A. (2005). Advanced treatment of benzothiazole contaminated waters: comparison of O3, AC, and O3/AC processes. Wat. Sci. Technol., 52, 281-288. Vitzthum, O.G., Werkhoff, P. & Hubert, P. (1975). New volatile constituents of black tea aroma. J. Agric. Food Chem., 23, 999-1003.

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Wallnöefer, P., Tillmanns, G., Thomas, R., Wünsche, C., Kurz, J. & Jarczyk, H.J. (1976). Mikrobieller Abbau des Herbizids Methabenzthiazuron und Identifizierung der Metaboliten. Chemosphere, 5, 377-382. Wegler, R. & Eue, L. (1977). Chemie der Pflanzenschutz- und SchädlingsBekämpfungsmittel. 5. Herbizide. Wegler R. (ed). Springer-Verlag, Berlin-HeidelbergNew York, 752p. Whittaker, M.H., Gebhart, A.M., Miller, T.C. & Hammer, F. (2004). Human health risk assessment of 2-mercaptobenzothiazole in drinking water. Toxicol. Ind. Health, 20, 149163.

In: Environmental Biodegradation Research Focus Editor: B. Y. Wang, pp. 189-215

ISBN: 978-1-60021-904-7 © 2007 Nova Science Publishers, Inc.

Chapter 7

BIODEGRADABLE ALIPHATIC POLYESTERS DERIVED FROM 1,3-PROPANEDIOL: CURRENT STATUS AND PROMISES George Z. Papageorgiou and Dimitrios N. Bikiaris♦ Laboratory of Organic Chemical Technology, Department of Chemistry, Aristotle University of Thessaloniki, Macedonia, Greece.

ABSTRACT Among biodegradable polymers, polyesters derived from aliphatic dicarboxylic acids and diols are of special importance. Polyesters of 1,3-propanediol were overlooked till recently, since the specific monomer was not available in the quantities and price that might enable production of polymers. However, in recent years more attractive processes have been developed for the production of 1,3-propanediol from renewable resources. Nowadays, research on biodegradable poly(1,3-propylene alkanedioate)s, such as poly(propylene succinate) (PPSu), poly(propylene adipate) (PPAd) and poly(propylene sebacate) (PPSe), has gained increasing interest, due to their fast biodegradation rates and their potential uses in biomedical or pharmaceutical applications, such as drug delivery systems. The odd number of methylene units in the diol segment is responsible for the lower melting points, lower degree of crystallinity and higher biodegradation rates of the specific polymers compared with their homologues based on ethylene-glycol or 1,4butanediol. In this chapter synthesis and properties of the 1,3-propanediol based aliphatic polyesters and especially their biodegradation characteristics are reviewed. Specific attention has been paid to preparation of related copolymers and blends with other important polymers, since these techniques may offer routes for optimizing properties and producing tailor-made materials. Copolymerization of 1,3-propanediol with mixtures of aliphatic or even aromatic acids, leads to linear polyesters with improved or balanced ♦Send correspondence to Dimitrios N. Bikiaris, Laboratory of Organic Chemical Technology, Department of Chemistry, Aristotle University of Thessaloniki, 541 24 Thessaloniki, Macedonia, Greece; Tel.: +30 2310 997812; Fax: +30 2310 997769; E-mail address: [email protected]

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1. INTRODUCTION Polyesters of 1,3-propanediol only recently were studied, since the specific diol monomer was not available in the quantities and price that might enable production of polymers. Since in recent years more attractive processes have been developed for the production of 1,3propanediol (1,3-PD) from renewable resources, research on related polymers has attracted interest from both an industrial and academic point of view [1]. Poly(propylene terephthalate) (PPT), the first and most studied polyester of 1,3propanediol, is available in the market [2]. The polymer is suitable for industrial fiber production. PPT fibers are characterized by much better resilience and stress/recovery properties than PET and PBT. These properties are due to the crystal structure of PPT. PPT chains are much more angular structured than PET and PBT chains due to the even number of methylene groups of the diol segment. Therefore these chains can be stretched up to 15% with a reversible recovery [3]. PPT is anticipated to gain a significant share in the thermoplastic polyesters market in the next years. However, like the other terephthalate polyesters, PPT is not susceptible to degradation in the environment. Recently, biodegradable aliphatic polyesters, and especially poly(propylene succinate) (PPSu) and poly(propylene adipate) (PPAd) have gained increasing interest [4-6]. Their synthesis, thermal properties and biodegradation were studied in comparison with the familiar polyesters poly(butylene succinate) (PBSu), poly(ethylene succinate) (PESu), poly(butylene adipate) (PBAd) and poly(ethylene adipate) (PEAd), which are also important biodegradable polymers. Poly(butylene terephthalate-co-adipate) copolymers are industrially produced in an attempt to arrive at polymers having both the advantages of high performance and being susceptible to degradation in environmental conditions [4]. Furthermore, high molecular weight PBSu is already available in the market. The odd number of methylene units in the diol segment is responsible for the lower melting points, lower degree of crystallinity and higher biodegradation rates of the polymers prepared form 1,3-propanediol, compared with their homologues based on ethylene-glycol or 1,4-butanediol [5]. In the following chapter an effort is made to summarize results of such works dealing with biodegradable polyesters of 1,3-propanediol.

2. DISCUSSION 2.1. 1,3-Propanediol as a monomer for polymer production 1,3-Propanediol became available in the market in sufficient quantity and purity, only a few years ago. Nowadays, more attractive processes have been developed for its production such as selective hydration of acrolein, followed by catalytic hydrogenation of the

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intermediate 3-hydroxypropionaldehyde or hydroformylation of ethyleneoxide (Figure 1). Recently, Du Pont and Shell Chemical companies have succeeded in producing 1,3propanediol commercial products with low cost and high quality by using different methods. [1]. 1,3-PD is a valuate chemical intermediate that has recently found extended applications as monomer for the production of polyesters [2, 7]. It is considered to be one of the bulk chemicals that will be produced in large scales in the future.

Figure 1. Industrial processes for synthesis of 1,3-propanediol.

Certain polymers like PPSu, PPAd and poly(propylene sebacate) (PPSe) beyond biodegradability, present the additional advantage that they can be produced from oligomers arriving from renewable sources. The production of chemicals and fuels by using alternative sources instead of hydrocarbons (oil), has received great attention as a green feed stock manufacture. 1,3-Propanediol, succinic, adipic and sebacic acid that are used as monomers for the preparation of biodegradable polyesters, are chemicals produced by using green techniques [4]. Until now 1,3-PD has been manufactured mainly by chemical synthesis, which requires high temperature, high pressure and expensive catalysts. Thus, in the last decade much effort has been paid to its production by bioprocesses on large scales. Many microorganisms like Klebsiella, Citrobacters, Enterobacter, Lactobacillus and Clostridia are able to convert glycerol or glucose to 1,3-PD via fermentation processes [8-11]. Experimental investigations showed that the fermentation is a complex bioprocess, taking place mostly in two stages while microbial growth is subjected to multiple inhibitions of substrate and byproducts. On the other hand, succinic acid is a dicarboxylic acid used as a monomer for the preparation of aliphatic and fully biodegradable polyesters. It can be produced

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pertochemically from butane and also by anaerobic fermentation, requiring glucose and CO2, decreasing the pollution caused from the petrochemical process. Glycerol and wood hydrolysates can be also used as carbon sources for succinic acid production, during fermentation by Mannheimia succiniciproducens bacterium [12,13].

2.2. Synthesis and characterization of the polyesters of 1,3-PD The synthesis of aliphatic polyesters with high molecular weight, in order to achieve satisfactory mechanical properties, is considered as being one of the most difficult problems to be solved. Till today this can be achieved only by either using techniques such as ringopening polymerization of cyclic monomers (lactones) or with the use of chlorides of acids, which are very expensive and inappropriate for industrial scale use [14,15]. The production of high molecular polyesters using diacids and diols can proceed only by the addition of chain extenders or branched comonomers as is the case of Bionolle® [16]. Synthesis of the aliphatic polyester samples from 1,3-PD was performed following the two-stage melt polycondensation method (esterification and polycondensation) in a glass batch reactor [17]. The synthesis is almost identical to that used for the preparation of PET with the only difference the lower temperature used for the polycondensation step. In brief, the proper amount of diacid (e.g. succinic acid or adipic acid) (0.55 mol) and 1,3-propanediol in a typical molar ratio 1/1.2 and the catalyst tetrabutoxy titanium TBT (typically 3×10-4 mol TBT/mol of diacid) were charged into the reaction tube (250 ml) of the polycondensation apparatus. The reaction mixture was heated at 190oC in an argon atmosphere until the collection of almost all the theoretical amount of H2O. In the second step of polycondensation, polyphosphoric acid (PPA) was added (5×10-4 mol PPA/mol of diacid) and a vacuum (5.0 Pa) was applied slowly over a time period of about 30 min. For each polyester, the polycondensation temperature was kept stable at 230oC while stirring speed was slowly increased to 720 rpm. The polycondensation reaction was finished after 2h of heating. From the studies it was found that monomer ratio, initial produced oligomers and polycondensation temperature are crucial parameters, which should be taken into consideration in order to obtain high molecular weight polyesters [18]. As reported above, the polymerization process involves two different steps according to the well-known process used for polyesters synthesis. In the first stage (esterification), aliphatic diacid reacts with 1,3-propanediol and water is removed as by-product. At the temperature that reaction takes place (190oC), water can be easily removed by distillation from the reactor and oligomers are formed. At this stage it is very important to remove as much of the formed water as possible and oligomers with the highest possible molecular weight to be produced. In order to increase the molecular weight at the appropriate level in the second stage (polycondensation), the prepared oligomers condensate at high temperature (230oC) with the application of high vacuum. This temperature is probably the maximum possible for the preparation of high molecular weight aliphatic polyesters, as was concluded from our study [19,20]. Above this temperature decomposition reactions can take place, which despite their very slow rate can reduce the molecular weight of the final polyester and color the sample. The reactions that occur during these stages (esterification and polycondensation) and the procedures used for the synthesis of the 1,3-PD polyesters are presented in Figure 2.

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Figure 2. Synthesis of poly(propylene alkanedioate)s following the polycondensation method.

Using different aliphatic diacids with increasing number of methylene units (X), from succinic with two methylene units to sebacic acid with eight methylene units, a series of polyesters of 1,3-propanediol was prepared. Thus, poly(propylene succinate) (X=2), poly(propylene glutarate) (X=3), poly(propylene adipate) (X=4), poly(propylene pimelate) (X=5), poly(propylene suberate) (X=6), poly(propylene azelate) (X=7) and poly(propylene sebacate) (X=8) samples were synthesized. Bikiaris and Achilias, following the polycondensation method and with varying the catalyst amount and/or the temperature of the second step of the process, prepared a series of PPSu samples with different molecular weight values [18]. In general it was found that with increasing the catalyst/diacid molar ratio in the reaction mixture, higher average molecular weight could be achieved, for given polycondensation temperature. Also, polymers with higher number average molecular weight (Mn) values were prepared by increasing the temperature of polycondensation up to 230oC. Finally, from theoretical simulation results it was found that although higher initial ratios of glycol to succinic acid are useful to increase the esterification rate, they lower the number average degree of polymerization of the oligomers at a fixed conversion of acid end groups. The 1HNMR spectra of PPSu and PPAd were reported [5,6]. In the 1HNMR spectra for PPSu (Figure 3) a single peak appeared at 2.63 ppm which was attributed to methylene protons of succinic acid, c, and a triple peak 4.09-4.21 ppm attributed to b protons and a multiple peak between 1.9-2.02 ppm corresponding to a protons were also observed [5]. In the spectra of PPAd, also shown in Figure 3, peaks for the a and b protons of 1,3-propanediol are recorded as for PPSu, while peaks for d and e protons are also observed at 1.67ppm and 2.36ppm respectively [6].

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PPSu

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Figure 3. Chemical structures and 1HNMR spectra of PPSu and PPAd.

The WAXD patterns of the prepared series of polymers, from poly(propylene succinate) (X=2) to poly(propylene sebacate) (X=8), are quite different as one can see in Figure 4. In general the parameters of the unit cell characterizing the crystals of the polymers have not been reported. However, the crystal structure of poly(propylene sebacate) (PPSe) has been studied [21]. PPSe has an orthorhombic unit cell of dimensions a=0.532, b=0.7532 and c=3.133nm (fiber axis) and belongs to the P212121 space group. There are two chemical units per fiber repeat and the two chains within the unit cell are positioned on 21 screw axis, parallel to the c direction. The fiber repeat is 0.25nm shorter than that for an all-trans conformation. The non-trans torsion angles are located in the glycolic moiety of the chemical repeat. It is anticipated in general the unit cells for polyesters of 1,3-PD to be orthorhombic in contrast to the butylene homologues for which monoclinic unit cells have been reported [21].

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PPSebacate PPAzelate

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The melting point variation for these polyesters with increasing methylene units in the diacid moiety, in comparison to the corresponding for 1,2-ethanediol and 1,4-butanediol can be seen in Figure 5. To construct the plots, data from the paper of Müller et al were also used [4]. It is obvious that for the 1,3-propanediol polyesters the variation of the melting points is small compared to those polyesters of the other two diols. 200

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It was observed that the crystallization rates of the polyesters of 1,3-propanediol in general increased with increasing the number of methylene groups in the respective diacid. Poly(propylene succinate) is a slowly crystallizing polyester which shows a glass transition temperature Tg at about -35oC and a melting temperature Tm=46oC. The latter is much lower than the corresponding values for PESu (Tm=104oC) or PBSu (Tm=112oC), as a result of the well-known odd-even effect for the melting temperatures of polyesters. On the other hand the Tg value is intermediate to those for PESu (-10oC) and PBSu (-43oC). The polymer does not crystallize during DSC cooling scans from the melt or on heating from the glassy state, unless very slow rates such as 2.5oC/min or even slower, are applied [20]. Figure 6a shows the respective heating scans for PPSu after melt-quenching. Poly(propylene adipate) crystallizes faster than PPSu due to the increased number of methylene groups in its repeating unit [6]. This is obvious in the DSC traces of Figure 6b, where the behavior of both quenched PPSu and PPAd on heating from the amorphous state is presented. In contrast to PPSu, PPAd shows cold-crystallization on heating by 10oC/min. Also, it has a Tg at –57oC, much lower than that of PPSu and a melting point of about 40oC for well-crystallized samples or lower for coldcrystallization ones. The values for the degree of crystallinity that the two polymers usually achieve are comparable, ranging close to 35% for both. For PPSu similar to the other two important succinates i.e. poly(ethylene succinate) (PESu) and poly(butylene succinate) (PBSu), multiple melting behavior was observed [20]. As can be seen in Figure 6a, double melting peaks are observed after cold crystallization. This behavior however, was more clearly shown for samples after isothermal crystallization from the melt. For the interpretation of these observations the partial melting-recrystallization-final melting scheme was adopted. Multiple peaks were more clearly revealed in DSC traces after crystallization in the temperature range from 0 to 20oC. The equilibrium melting temperature (Tmo) of PPSu and the heat of fusion for the pure crystalline polymer were estimated to be 58oC and 22kJ/mol respectively. This Tmo value compared to 114oC and 133.5oC for the Tmo of PESu and PBSu respectively, is quite lower.

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Figure 6. a) DSC heating traces for quenched PPSu at different slow heating rates and b) DSC traces of amorphous PPSu and PPAd recirded at 10oC/min.

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Dielectric spectra measurements, upon crystallization of PPSu, have been performed by Soccio et al [22]. Crystallization was performed at 25oC. After a controlled period of time the sample was rapidly cooled to –25oC at 5oC/min to perform a frequency swept. PPSu at –25oC that is a temperature above the glass transition temperature exhibits two main relaxations processes characteristic of most polymers, α and β, in order of increasing frequency. It was found that for the polyester the a and β relaxations appear simultaneously and are well resolved in the experimental frequency window. The β relaxation was used to characterize the crystalline structural development while the a relaxation was used to collect information about the evolution of the amorphous phase dynamics. In this way structural development could be characterized by a single experiment during the crystallization process. The analysis of the dielectric loss supported the existence of precursors of crystallization in the induction period.

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The thermal degradation behavior of the polyesters has been studied [6,23,24]. From the thermogravimetric curve of Figure 7 it can be seen that PPSu presents a relatively good thermostability since no significant weight loss (only1.2 %) occurred until 300oC. As it can be seen in the curve of derivative weight loss curve (DTG), in the early stages of the decomposition, there is a small shoulder peak. Until 366 oC, where the maximum rate of this decomposition step appears, the volatile matter corresponds to about 7 wt% of initial weight. Such a pre-major weight loss stage was also mentioned for poly(propylene terephthalate) (PPT) with low number average molecular weights ranging between 13000 and 23000 g/mol, and PCL [25-27]. The temperature at the maximum weight-loss rate of this stage increases significantly with molecular weight while the weight loss decreases steadily. The first decomposition step (shoulder) is attributed to degradation and volatilisation of the oligomers detected with 1H-NMR spectroscopy. From preliminary studies of the monomers it was found that succinic acid degrades at temperatures close to 200oC while 1,3-propanediol at slightly higher temperatures. However, both are fully decomposed at temperatures lower than 300oC. These data verified the hypothesis that the first decomposition step is due to oligomer

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degradation. The temperature at which PPSu decomposition gains the highest rate is at 404oC for heating rate 10 oC/min (Figure 7). This temperature is very high taking into account that is aliphatic polyester and the temperature value is comparable to decomposition temperatures reported for aromatic polyesters like terephthalates (PET, PBT, PPT) and naphthalates (PEN) [17]. Furthermore, this temperature is much higher compared to other aliphatic polyesters like poly(L-lactide) for which major degradation step appears at temperatures lower than 370oC, as well as poly[(R)-3-hydroxybutyrate] and poly(ε-caprolactone) [28, 29]. From the above it can be concluded that PPSu although it has low melting point presents a very high stability against thermal degradation. Thermal degradation of PPAd was also studied by determining its mass loss during heating [6]. In Figure 7 the remaining mass (TG %) and the derivative (DTG) curves at heating rate 10oC/min are presented for the polyester. In both TG and DTG thermograms one stage of mass loss is followed and as can be seen the weight loss takes place gradually and smoothly. Additionally, there is no differentiation between using low or high heating rates. PPAd presents a relatively good thermostability since no significant weight loss, ~0.25%, occurred until 250oC. After that temperature the polyester decomposes quickly and looses about 95.8 wt% until 500oC. The temperature at which PPAd decomposition gains the highest rate is 385oC, for a heating rate of 10oC/min. However, the behaviour of the polyester is not similar to that of poly(butylene adipate) (PBAd) and poly(propylene succinate) (PPSu) and has a lower thermal stability. Finally, the decomposition temperature of PPAd is 20-25oC lower than the corresponding temperature of PPSu. The mechanical properties of PPAd were also reported [6]. It has a tensile strength of 11.6 MPa, which is similar with that of low density polyethylene 13.2 MPa (LDPE). However its homologues PEAd and PBAd have much higher tensile strength, reaching 17.3 and 16.2 MPa, respectively. Furthermore, for high molecular weight PPAd samples satisfactory elongation at break was observed, higher than 400%. PPSu on the other hand has a very low tensile strength 1.6 up to 3.8 MPa depending on its molecular weight and breaks before yielding [5].

2.3. Biodegradation Biodegradability of a certain polymer in the form of enzymatic hydrolysis is controlled by several factors. The most important one is the nature of the polymer itself, meaning its chemical structure and the occurrence of specific bonds, which might be susceptible to hydrolysis, along its chains. Such groups are those of esters, ethers, amides, etc. [30,31]. Enzymatic hydrolysis studies of PPSu have been carried out [5]. For comparison, the behavior of PESu and PBSu samples having the same molecular weight values, (number average molecular weight of about 6800g/mol), were also studied, at the same conditions. The polyesters in the form of films of 5x5 cm in size and approximately 2 mm thickness, prepared in a hydraulic press, were placed in petries containing phosphate buffer solution (pH 7.2) with 1 mg/mL Rhizopus delemar lipase. Usually enzymatic hydrolysis studies are performed at 37oC. However this temperature is very close to the melting point of PPSu. So, in this case the tests were performed at 30±1oC. The degree of biodegradation was estimated from the mass loss and molecular weight reduction as measured by GPC. The results showed that PPSu exhibits the highest degradation rates (Figure 8a), while PBSu the lowest ones as one

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Remaining weight (mg/cm2)

can see in [5]. For PESu weight loss rate seemed to be similar to that of PBSu, but final mass loss was slightly larger. Since the used polyester samples had almost identical molecular weights, crystallinity seems to be the predominant factor that controls the biodegradation rates. It is therefore the higher degree of crystallinity of the PBSu samples, as was found by DSC and WAXD, which resulted in lower degrees of biodegradation comparing to that of PESu or PPSu. PPSu, which shows the lower crystallinity and melting point, hydrolyzes faster comparing to the other polyesters.

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Figure 8. Variation of the weight of specimens of a) PPSu, PESu and PBSu during enzymatic hydrolysis using Rhizopus delemar lipase and b) PPAd, PEAd and PBSu during enzymatic hydrolysis using 0.09 mg/mL Rhizopus delemar lipase and 0.01 mg/mL of Pseudomonas Cepacia lipase.

In recent studies the biodegradation characteristics of PPSu, PPAd and PPSe, as well as some relative copolymers were reported [32, 33]. The biodegradability of the polyesters was detemined by monitoring the normalized weight loss of polyester films with time in phosphate buffer (pH=7.2) without and with Rhizopus delemar lipase at 37oC. It was found that the biodegradation rates of the homopolyesters follow the path PPSu>PPAd>PPSe. Poly(propylene sebacate) did not show significant weight loss in presence of enzyme which may be due to its highest degree of crystallinity and melting point compared to PPSu, PPAd and copolyesters. The PPSu and PPAd copolyesters with PPSe showed slower degradation rates than the PPSu and PPAd homopolyester respectively. However, from recent work we have concluded that PPAd degrades faster than the other adipate polyesters, PEAd and PBAd, as can be seen from Figure 8b. At very sort time, only 6 days, the whole amount of PPAd has

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been enzymatically hydrolyzed, while at the same time only the 40 wt% of PBAd has been hydrolyzed. SEM microphotographs revealed specific features indicative for surface erosion in the case of PPSu as can be seen in Figure 9. In general the observations lead to conclusions consistent with those derived from the weight loss and molecular weight measurements during the experiments. As it is well known enzymatic hydrolysis is a heterogeneous process and enzymes are attached on the surface of an insoluble substrate and hydrolysis takes place via surface erosion. During this procedure small fragments and water soluble monomers or oligomers are generated, during hydrolysis of ester bonds. In general, the interior of polyesters specimens is not attacked until extended holes are created on the surface allowing the enzymes to enter and attack the main body. After consumption of the amorphous material of the surface a layer of crystalline domains remains, where only slow degradation may occur. As can be seen in Figure 9 the crystallites of PPSu are well distinguished after 35 days of enzymatic hydrolysis and remain almost unaffected. In principle, hydrolysis affected the amorphous material surrounding the spherulites and numerous holes were then created.

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b

Figure 9. SEM micrographs of PPSu polyester during hydrolysis at a) 15 days and b) 35 days.

2.4. Copolymers Recently, except from homopolymers, copolymers based on 1,3-propanediol, succinic acid and other diacids or diols were synthesized [34]. A full series of random poly(butyleneco-propylene succinate) (PBPSu) copolymers with high molecular weight were prepared, by the polycondensation method. The intrinsic viscosity of the copolymers ranged from 0.6 up to 0.75 dL/g, but without showing any clear trend for the dependence on composition. The copolyesters showed a melting point depression with increasing comonomer content, as can be seen in Figure 10. Also, on the basis of WAXD observations and the decrease in melting points, isodimorphic cocrystallization was concluded. The probability, comonomer propylene succinate units to be incorporated in the PBSu crystals, was higher for high comonomer content. As the comonomer content increases, the mean length of sequences of the same repeating unit along the macromolecular chains, are expected to reduce. This in turn results in formation of non-crystallizable chain segments, and finally the crystallinity decreases. In the case that comonomer units are inserted in the crystals, they act as crystal defects, lowering the

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melting point of the copolymers. The most important feature of these PBPSu copolymers however, was that of their fast enzymatic hydrolysis rates.

o

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Propylene succinate (mol%) Figure 10. Melting points and glass transition temperatures for the poly(butylene-co-propylene succinate) copolymers as a function of propylene succinate units content.

Figure 11 shows that the copolymers with high propylene succinate content experienced faster degradation rates, even than those of the neat PPSu. This fact was attributed to their lower degree of crystallinity, lower melting points and increased mobility in the amorphous phase, which is known that is general the phase that is mainly affected by enzymatic hydrolysis. 100

PPSu90 PPSu60 PPSu50 PPSu40 PPSu30 PPSu20 PPSu10 PPSu5 PBSu PPSu

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Enzymatic hydrolysis (days) Figure 11. Weight loss as a function of time of enzymatic degradation of PBPSu copolyesters compared to the neat PPSu and PBPSu.

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Figures 12 and 13 show SEM photographs of the surfaces of copolymer specimens taken after different times of enzymatic hydrolysis. Surface erosion is more pronounced for higher content of propylene succinate units in the copolymers.

a

b

Figure 12. SEM micrographs of PBPSu 70/30 w/w copolymer during enzymatic hydrolysis at several times a) 0 days and b) 15 days.

Figure 13. SEM micrographs of PBPSu 10/90 w/w copolymer during enzymatic hydrolysis at several times. From left to the right: 0 days, 3 days and 15 days.

Also, two PBPSu copolyesters were studied with respect to their isothermal crystallization kinetics [35]. The Avrami exponent was found to be between 2.2 and 2.8, showing that the crystallization mechanism was a heterogeneous nucleation with spherical growth geometry in the crystallization process of the polyesters. Multiple melting peaks were observed. PBPSu copolymers were identified to have the same crystal structure with that of PBSu by using WAXD. It was assumed that only butylene succinate units could crystallize, while the propylene succinate remain in the amorphous state. Other studies showed that this hypothesis might hold for low propylene succinate content [34]. Poly(propylene succinate-co-ε-caprolactone) PPSu/PCL copolymers with compositions 95/5, 90/10, 25/75, 50/50 and 75/25, w/w were also prepared according to the procedures described by Seretoudi et al., using a combination of polycondensation and ring opening polymerisation [36]. The degree of randomness of the copolymers was characterized by using 1 H-NMR and 13C NMR and was verified that at concentrations up to 25 wt% of PPSu, block copolymers were prepared while at compositions 50 and 75 wt% random copolymers instead of block were prepared [37]. This is due to the extensive transesterification reactions that are taking place simultaneously with the ring-opening polymerization of ε-CL. The molecular

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weight distribution of the prepared polyesters was determined by GPC and it was found that as the amount of PCL increases in the copolymer the molecular weight, also, increases. This is attributed to the fact that as the amount of initial PPSu increases, higher amounts of hydroxyl end groups contained in PPSu are available to act as initiators for ring opening polymerization of ε-CL monomer. However, despite the high molecular (Mw) of the copolymers, which ranged between 55000 and 87000 g/mol, their tensile strength is very low (1.3 up to 3.4 MPa) and only their elongation at break is higher than 560% for all copolymers. This behavior could be also attributed to their crystallinity. Random copolymers were completely amorphous in contrary to the block, which were semicrystalline.

Intensity (a.u.)

PPSu PCL/PPSu 95/5 PCL/PPSu 90/10 PCL/PPSu 75/25 PCL/PPSu 50/50 PCL/PPSu 25/75 PCL

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o

Diffraction angle 2θ ( ) Figure 14. XRD patterns of PPSu, PCL and their copolymers.

In DSC thermograms recorded during heating of the copolymers containing 50 and 75 wt% PPSu only single glass transitions could be observed, which are located at –53.8 and at – 43.9oC respectively. These temperatures are lower than the glass transition temperature of PPSu (–35oC) and higher than the Tg of PCL (-61.6oC). So, the Tgs of the prepared copolymers were between the Tgs of the neat polymers. For the other copolymers two separate melting peaks were observed close to the melting temperatures of the neat polyesters. This is an indication that the different blocks can crystallize separately. Thermal degradation of PPSu/PCL 25/75, 50/50, and 75/25 w/w copolymers was studied by determining their mass loss during heating. The prepared copolymers showed satisfactory stability against thermal degradation since no significant weight loss, (~0.5%), occurred until 250oC while their maximum decomposition rate occurred at temperatures about 410oC. The corresponding temperatures for PPSu and PCL are 403.9 and 415.3 oC, respectively. Concerning their thermal decomposition mechanism it is proved that all copolymers degrade at two stages. The first stage corresponded to small mass loss taking place at 280-365oC.

George Z. Papageorgiou and Dimitrios N. Bikiaris

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PPSu/PCL 25/75 w/w

PPSu/PCL 50/50 w/w

-60

0

60 o

Temperature ( C) Figure 15. DSC thermograms of PPSu/PCL copolymers.

The most characteristic behavior of these copolymers is their high enzymatic hydrolysis at phosphate buffer solution (pH 7.2, 30oC) using 0.09 mg/mL Rhizopus delemar lipase and 0.01 mg/mL of Pseudomonas Cepacia lipase. PCL it is well known that due to its high hydrophobicity and crystallinity has a slow degradation profile and after 9 days of enzymatic hydrolysis its weight loss was less than 5 wt%. On the contrary almost all the copolymers at the same time were completely degraded as can be seen from Figure 16. The rate of enzymatic hydrolysis was slightly dependent on the PPSu content and increases by increasing its amount in the copolymer. This behaviour should be attributed to the higher biodegradation rate that PPSu has compared with PCL and to the fact that the copolymers were amorphous. It is well known that first the amorphous phase is degraded during enzymatic hydrolysis, while the crystalline phase due to the low water absorption rate is much less affected. Soccio et al reported synthesis of random poly(propylene isophthalate-co-succinate) (PPISu) copolyesters [38]. The copolymers also showed depression in the melting temperature with comonomer content. It was found that they could crystallize at room temperature, except those with 70 or 80 mol% propylene succinate (PSu) units. For up to 75 mol% propylene succinate units in the copolymers, crystals of poly(propylene isophthalate) were formed, while the copolyester with 90 mol% PSu gave crystals of PPSu. TGA measurements showed high thermal stability for the copolymers. Initiation of decomposition for PPSu was found to occur at 393oC, whereas the maximum rate appeared at 427oC, during scans by 10oC/min in air. The respective temperatures increased for the copolymers with increasing isophthalate content.

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PCL/PPSu 75/25 w/w PCL/PPSu 50/50 w/w PCL/PPSu 25/75 w/w PCL

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Time of enzymatic hydrolysis (days) Figure 16. Weight loss of PCL/PPSu copolymers during enzymatic hydrolysis for several days.

Similar observations were reported for random poly(propylene isophthalate-co-adipate) (PPIAd) copolyesters [39]. For up to 60 mol% propylene adipate (PAd) units in the copolyesters they formed PPI crystals. The copolymers with 70 or 80 mol% propylene adipate could not crystallize at room temperature, while that with 90 mol% PAd units formed PPAd crystals. The thermal stability of the samples was checked by TGA measurements in air by heating scans by 10oC/min. The polyesters were found to be stable. For PPAd the temperatures for initialization of decomposition and that of maximum decomposition rate were reported to be 359 and 392oC respectively. The respective temperatures for the copolymers increased with increasing propylene isophthalate content. Albertsson et al have reported the synthesis of poly(ester urethane)s (PEU), and poly(ester carbonate)s prepared by chain extension [40-43]. High molecular weight poly(ester urethane)s were prepared via the diisocyanate synthesis (Figure 17). Oligo(propylene succinate)s with molecular weights of about 2300-2400 g/mol were chain extended after reaction with 4,4΄-diisophenylmethane diisocyante (MDI). In this way PEUs with 46 up to 63wt% aliphatic polyester were prepared. The Tg values ranged from –7.8 to 4.8oC, while the melting points were found to be between 176 and 208oC. The sample with the highest polyester content in the reactor feed showed the lowest melting point. The Young’s modulus decreased when the content of the aliphatic polyester increased.

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Figure 17. Synthesis of oligo(propylene succinate), by thermal polycondensation of 1,3-propanediol and succinic acid, and chain extension reaction by diisocyanate syntheis [42].

High molecular weight poly(ester carbonate)s were prepared via the dichloroformate synthesis [40]. α,ω-dihydroxy-terminated oligo(propylene succinate)s prepared by thermal polycondensation were chain extended using phosgene. The chain extension reaction was performed also in two steps (Figure 18). The oligomers were first dissolved in dry chloroform in the presence of acid acceptors like ethyldiisopropylamine. Reaction with excess phosgene resulted in α,ω-dichloroformate. The latter was then polycondensated with an equivalent amount of α,ω-dihydroxy-terminated oligo(propylene succinate)s to receive high molecular weight poly(ester carbonate)s. However the melting points and mechanical properties remained low even for high molecular weight poly(ester carbonate)s. For this reason chain extension of oligomers obtained by the esterification of succinic acid in the presence of 1,3propanediol and 1,4-cyclohexanedimethanol was also performed [43]. The resulting polymers were characterized by glass transition temperatures and melting points, which increased with increasing the 1,4-cyclohexanedimethylene succinate units content.

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Figure 18. Chain extension reaction by dichloroformate synthesis of oligo(propylene succinate) [40].

Finally, segmented copolymers based on oligo(propylene succinate)s and poly(ethylene glycol) were synthesized [43]. These showed a higher affinity towards polar solvents than oligo(propylene succinate)s and their water absorption capability was increased to 10-15wt%.

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Another strategy to prepare polyesters is from lactones (cyclic esters) [43]. There has been established a method for the preparation of 1,4-dioxepan-2-one from 1,3-propnadiol after reaction with chloroacetic acid sodium salt. Then, ring-opening polymerization in the presence of catalysts leads to polymers with high molecular weight (Figure 19).

Figure 19. Ring-opening polymerization of 1,4-dioxepan-2-one [43].

As a matter of fact aliphatic polyesters of 1,3-propanediol have rather low thermal and mechanical properties, compared to aromatic polyesters. Copolymerization of 1,3propanediol with aliphatic and aromatic acids might enable preparation of materials having both biodegradability and sufficient properties. Terephthalate polyesters such as poly(ethylene terephthalate) (PET), poly(propylene terephthalate) (PPT) and poly(butylene terephthalate) (PBT) are very important for industry, due to their high thermal and mechanical properties and finally their many applications. Although poly(propylene terephthalate) is a new polyester, it has been used in many applications already, especially in the form of fibers. It has a meting point of about 230oC, a glass transition temperature 44oC and crystallizes faster than PET, but slower than PBT. There is an increasing interest about this polymer and its applications are expanding. Terephthalates and in general aromatic polyesters cannot be degraded in the environment [44]. However, Witt et al. in relatively recent works, prepared oligo esters of terephthalic acid with 1,2-ethanediol, 1,3-propanediol, and 1,4-butanediol and these oligomers were investigated with regard to their biodegradability in different biological environments [4,45]. Well-characterized oligomers with weight-average molar masses of from 600 to 2600 g/mol exhibit biodegradation in aqueous systems, soil, and compost at 60°C. Size exclusion chromatography (SEC) investigations showed a fast biological degradation of the oligomer fraction consisting of 1 or 2 repeating units, independent of the diol component used for polycondensation, while polyester oligomers with degrees of polymerization higher than 2 were stable against microbial attack at room temperature in a time frame of 2 months. At 60°C in a compost environment chemical hydrolysis also degrades chains longer than two repeating units, resulting in enhanced degradability of the oligomers. Metabolization of the monomers and the dimers as well by the microorganisms could be confirmed by comparing size exclusion chromatography (SEC) measurements and carbon balances from specific experiments. Finally it was concluded that based on these results degradation characteristics of potential oligomer intermediates resulting from a primary chain scission from copolyesters consisting of aromatic and aliphatic dicarbonic acids can be predicted depending on their composition. These results will have an evident influence on the evaluation of the biodegradability of commercially interesting copolyesters and lead to new ways of tailor-made designing of new biodegradable materials as well.

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2.5. Blends Poly(propylene succinate) and poly(ethylene succinate) (PESu) are both biodegradable polymers. PPSu shows much faster enzymatic hydrolysis rates than PESu. On the other hand PESu shows better thermal and mechanical properties. Blending is an alternative to prepare tailor-made materials. Thus, blends of poly(propylene succinate) and poly(ethylene succinate) were prepared in order to explore their miscibility [46]. It was thought that, due to their similar chemical structures the polymers might show miscibility or compatibility. Indeed, they were found to be miscible. For the preparation of the PESu/PPSu blends polymers with the same intrinsic viscosity, about 0.28 dL/g were used. The number average molecular weight as determined by GPC, was close to 7000g/mol. These binary blends were prepared in chloroform as solvent. The solution of both polymers (0.0125g/ml) was cast on a petri dish at room temperature. The solvent was allowed to evaporate in air for 3 days. The compositions of the blends were PESu/PPSu 80/20, 60/40, 50/50, 40/60, 80/20 in weight ratio. As it was observed in the respective DSC traces, the blends showed single composition-dependent glass transition temperature (Tg) between the Tg values of the neat PESu and that of the neat PPSu. Usually, the glass transition temperature criterion can be safely applied if the Tgs of the components of the polymer blends differ by 15-20oC. The glass transition temperatures of PPSu and PESu differ by 24oC. Furthermore, for better resolution slow DSC scans for quenched samples were also preformed, by slow heating at 2.5 or 1oC/min. These traces also showed single glass transition. The evidence of single composition-dependent glass transition between those of the neat components for the PESu/PPSu blends showed that they are probably miscible over the entire range of composition. The variation of the Tg values of the blends was found that could be described by the Gordon-Taylor equation. The crystallization of the blends was studied using DSC. Though the neat PPSu, if amorphous, cannot show cold-crystallization the blends with less than 40wt% PPSu content showed coldcrystallization during scans by 20oC/min. Furthermore, for the specific blends after crystallization at low temperature, melting peaks for both components were observed. This showed that they could crystallize in the blends. In fact it was found that PESu crystallizes first, while PPSu crystallizes at a much lower temperature, in the confined environment, due to the presence of PESu crystals. Since PPSu is still liquid when PESu crystallizes, there is a melting point depression of the higher melting temperature component, PESu. Its melting point in the blends decreased steadily with increasing PPSu content. However, similar conclusions were derived for PPSu, when the blends crystallized at low temperature. The equilibrium melting points were determined for the neat polymers and the blends and the melting point depression for the high melting temperature component, PESu, was analyzed applying the Nishi-Wang theory. After that analysis the calculated values for the interaction parameter were found to be negative. This was also evidence that the specific blends were miscible in the melt phase. WAXD showed that the polymers could form their own crystals in the blends, since both diffractions associated with the PESu and the PPSu crystals were revealed in the blend patterns. However, the relative intensity of the peaks varied with blend composition. The fact that both components crystallized separately was attributed to the significant difference in the crystallization rates of the two polymers. The specific blends were classified as semicrystalline/semicrystalline. In contrast to blends between semicrystalline and amorphous polymers, which have been extensively studied, blends in which both components are semicrystalline and especially in which both are biodegradable

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Normalized Heat Flow (W/g)

Normalized Heat Flow (W/g)

are not often reported in the literature. In general it is believed, that such blends could be used alternatively to copolymers for pharmaceutical applications.

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b

Figure 20. DSC traces for amorphous and crystallized PESu/PPSu blends.

In contrast to PPSu which is a new polyester, PCL is well-known and extensively studied polymer. It is also biocompatible and has found many applications in pharmaceutical technology and medicine. Thus, it was also an interesting idea to explore miscibility and biodegradation behavior of blends made of these two very important polyesters. PCL/PPSu blends with concentrations 90/10, 80/20, 70/30 and 60/40 w/w were prepared by solutioncasting [47]. Proper amounts of both polymers were dissolved in chloroform as common solvent, at room temperature. Sonication was applied in order to achieve complete dissolution and fine mixing of the components. The blends in the form of thin films (200-250 μm) were set up after solvent evaporation at room temperature, under a gentle air stream. They were characterized by DSC, WAXD, 1HNMR, SEM, and Tensile testing. Finally, their enzymatic hydrolysis was studied. The PCL/PPSu blend system however proved to be only partially miscible. DSC studies on the melting point depression showed that there was a slight decrease in the melting point of the high temperature component, PCL, in the blends. Results after the analysis of the melting point depression using the Nishi-Wang model were marginal most probably showing a limited miscibility. Tensile properties testing however proved that the blends are at least compatible if not miscible. This was concluded because despite the decrease in the tensile properties of the blends with increasing PPSu content, there was no indication for any minimum. Crystallization rates of PCL were retarded with increasing second component content in the blends. Biodegradation of the blends was also explored. Enzymatic hydrolysis for several days of the prepared blends was performed using Rhizopus delemar lipase at pH 7.2 and 30oC. SEM micrographs of thin film surfaces revealed that hydrolysis affected mainly the PPSu polymer as well as the amorphous phase of PCL. For all polymer blends an increase in the values of the melting temperature and the heat of fusion

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was recorded after the hydrolysis. This was associated with the fact that enzymatic hydrolysis affected mainly the amorphous phase of the blends. The consumption of the amorphous portion of the specimens resulted in an increase of the estimated crystallinity after hydrolysis. The biodegradation rates as expressed in terms of weight loss were faster for the blends with higher PPSu content (Figure 21). PPSu was proved to have much higher biodegradation rates compared to PCL, due to the lower crystallinity, the decreased melting point (Tm for PPSu is about 45oC) comparing to PCL (Tm about 60-65oC), and the number of the ester bonds along the macromolecular chains for given molecular weight. SEM observations lead to conclusions aligned with those from the weight loss measurements. That is, only for blends with high PPSu content, extended surface erosion occurred even from the first days of enzymatic hydrolysis.

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Normalized weight loss (mg/cm )

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PCL PCL/PPSu 90/10 PCL/PPSu 80/20 PCL/PPSu 70/30 PCL/PPSu 60/40 PCL/PPSu 20/80 PPSu

18 16 14 12 10 8 6 4 2 0 -2 -5

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Enzymatic Hydrolysis Time (days) Figure 21. Weight loss of PCL/PPSu blends during enzymatic hydrolysis for several days.

Poly(propylene adipate) (PPAd) was reported to be miscible with poly(ethylene oxide) PEO. In fact Lin and Woo studied the miscibility of PEO with a series of polyesters [48]. They found that PEO is miscible with polyesters such as poly(ethylene adipate), poly(propylene adipate), poly(butylene adipate), and poly(ethylene azelate). On the other hand PEO was found to be immiscible with poly(1,6-hexamethylene adipate) and poly(1,6hexamethylene sebacate). Optical observations showed that the melts of the blends of PEO with the last two polyesters were cloudy and heterogeneous in contrast to all the others. Also, the Nishi-Wang model was used to analyze the melting point depression observed for PEO in the blends. The conclusion was that the ratio CH2/CO in the polyesters is crucial for the miscibility with PEO. PEO was miscible with polyesters with a CH2/CO ratio from 3 up to 4.5, while miscibility was favored especially in the case of a value 3.5, which is the case of PPSu.

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2.6. Application of PPSu in drug delivery systems Aliphatic polyesters can be used in pharmaceutical applications, like drug delivery systems. As was reported above PPSu has a low melting temperature (40 to 46oC depending on the molecular weight), slightly above the temperature of the human body. Consequently, PPSu might be an alternative for the preparation of solid dispersions of drugs. Solid dispersions are dosage forms whereby the drug is dispersed in a biologically inert matrix. They can be used for the increase of the dissolution rates and thus the bioavailability of sparingly water-soluble drugs. As a matter of fact an increasing number of poorly watersoluble drugs are synthesized, and thus there is a demand for finding routes to increase their solubility. Alternatively solid dispersions can be used for controlled release of drugs. In a recent research, four PPSu samples with intrinsic viscosities 0.1, 0.18, 0.28 and 0.50 dL/g measured in chloroform solutions, were used in the preparation of solid dispersions of the drug Fluvastatin, with a drug load 30 wt%. A second series of solid dispersions were prepared, where the PPSu 0.28 was used, while the drug load varied from 10 to 50 wt%. The solid dispersions were prepared by the melt method, which involves dissolving the drug in the molten polymer, and then cooling to the room temperature. WAXD patterns of the solid dispersions proved that the drug was amorphous in all cases, since only the peaks corresponding to the crystalline polymer were revealed (Figure 22).

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Figure 22. WAXD patterns of a) Fluvastatin solid dispersions in PPSu 0.28, with different drug loadings and b) Fluvastatin solid dispersions in PPSu samples of different molecular weight and containg 30wt% drug.

The release rates of the drug from the solid dispersions increased with decreasing the molecular weight of PPSu, probably because low molecular weight polymers are brittle and increased porosity allows drug to come in contact to the aquatic dissolution medium (Figure 23). Also, for given molecular weight of PPSu the rates increased with increasing the drug content in the solid dispersions.

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Figure 23. Release of Fluvastatin from solid dispersions in PPSu during in-vitro tests: a) solid dispersions prepared using PPSu matrices with different molecular weights and constant drug load (30wt%), and b) solid dispersions prepared using the PPSu 0.28 sample and different drug loads.

3. CONCLUSION New industrial processes allowed production of 1,3-propanediol, a very important monomer for the synthesis of polyesters. Biodegradable polyesters based on 1,3-propanediol have gained an increasing interest. Especially polymers like poly(propylene succinate) are fully biodegradable and can also be produced from monomers from renewable resources. The polyesters can be synthesized following the polycondensation method. Recent experimental work on enzymatic hydrolysis has shown that polyesters of 1,3-propanediol have faster degradation rates compared to the corresponding from ethylene glycol or 1,4-butanediol. Poly(propylene succinate) seems to be faster degraded than poly(propylene adipate) or poly(propylene sebacate). Copolymers prepared from polycondensation of succinic acid with 1,3-propanediol and 1,4-buanediol showed improved biodegradation characteristics compared even to the neat PPSu. Also, copolymers synthesized from oligo(propylene succinate) and εcaprolactone, combined characteristics of the homopolymers, showing improved hydrolysis behavior. Combination of succinic or adipic acid with isophthalic acid as starting materials for polycondensation with 1,3-propanediol gave examples of copolymers containing both segments which can hydrolyze and aromatic ones in the macromolecular chains. Presence of aromatic moieties is desirable in order to increase strength of the materials. Alternatively, starting from oligo(propylene succinate)s, methods have also been developed to produce polymers like poly(ester urethane)s or poly(ester carbonate)s with high molecular weight, which finally are also biodegradable. Fully biodegradable blends of PPSu with PESu have been found to be miscible. Blends of PPSu with poly(ε-caprolactone) showed only very limited miscibility, but satisfactory tensile properties. In general, results from the research on 1,3-propanediol based polymers are very encouraging. Polyesters of this class are expected to find applications, especially in pharmaceutical technology, and their uses in such applications have already been tested. However, generally speaking and taking into account the parameter cost of biodegradable polymers, much work should be carried out to arrive at optimal solutions and wide applications.

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In: Environmental Biodegradation Research Focus Editor: B. Y. Wang, pp. 217-238

ISBN: 978-1-60021-904-7 © 2007 Nova Science Publishers, Inc.

Chapter 8

AEROBIC BIODEGRADATION OF FISH-MEAL WASTEWATER FROM LAB SCALE TO LARGE SCALE Joong Kyun Kim1 and Geon Lee2 1

Department of Biotechnology and Bioengineering, Pukyong National University, Busan, Korea 2 Research Department, Samrim Corporation, Kimhae, Kyongnam, Korea

INTRODUCTION The amount of fisheries waste generated in Korea is expected to increase with a steady increase in population to enjoy taste of slices of raw fish. The fisheries waste is reduced and reutilized through the fish meal production. Depending on the raw material used, there are basically two types of fish-meal manufacturing processes: those that use fish wastes, such as heads, bones or other residues, and those that use the whole fish. The process, using fish wastes, is the commonest used in the Korean industries. The first step of the fish-meal manufacturing processes is the compression and crushing of the raw material, which is then cooked with steam, and the liquid effluent is filtered off in a filter press. The solids obtained are introduced to a rotating drier and finally cut and crumbled to obtain the commercial fishmeal product. The liquid stream contains oils and a high content of organic suspended solids. After oil separation, the fish-meal wastewater (FMW) is generated with stinky odor and shipped to wastewater treatment place. FMW has been customarily disposed of by dumping into the sea, since direct discharge of FMW can cause serious environmental problems. Besides, bad smell, which is produced during fish-meal manufacturing processes, causes civil petition and stricter regulations for this problem come into force every year in Korea. Therefore, there is an urge to seek for an effective treatment to remove the organic load from the FMW; otherwise the fish meal factories will be forced to shut down. Biological treatment technologies of fish-processing wastewater have been studied to improve effluent quality (Battistoni and Fava, 1995; Park et al., 2001). The common feature of the wastewaters from fish processing is their diluted protein content, which after concentration by a suitable method would enable the recovery and reuse of this valuable raw

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material, either by direct recycling to the process or subsequent use in animal feed, human food, seasoning, etc. (Afonso and Borquez, 2002). It has been reported that the organic wastes contain compounds, which are capable of promoting plant growth (Day and Katterman, 1992), and seafood processing wastewaters do not contain known toxic or carcinogenic materials unlike other types of municipal and industrial effluents (Afonso and Borquez, 2002). Therefore, FMW could be a valuable resource for agriculture. However, potential utilization of this fish wastes has been limited because of its bad smell (Martin, 1999). There is an increasing need to find ecologically acceptable alternatives to overcome this problem. Aerobic biodegradation has been widely used in treatment of wastewaters, and recently references to the use of meso- and thermophilic microorganisms have become increasingly frequent (Cibis et al., 2006). During the biodegradation, the organic matter is biodegraded mainly through exothermic aerobic reactions, producing carbon dioxide, water, mineral salts, and a stable and humified organic material (Ferrer et al., 2001). There have been few reports that presented the reutilization of biodegraded waste products as liquid-fertilizer: a waste product of alcoholic fermentation of sugar beet (Agaur and Kadioglu, 1992) and diluted manure streams after biological treatment (Kalyuzhnyi et al., 1999). Therefore, aerobic biodegradation is considered to be the most suitable alternative to treat FMW and realize a market for such a waste as fertilizer. Moreover, the reutilization of FMW can create the high additional value, since FMW is collected from the industry with security for cost of FMW treatment. Scale-up is the study of the problem associated with transferring data obtained in laboratory equipment to industrial production. It is clear that problems of scale-up in a bioreactor are associated with the behavior of liquid in the bioreactor and the metabolic reactions of the organisms. Biological properties, especially various constants involved in kinetic equations depend on scale-up, although the metabolic patterns remain unchanged. The typical differences of bioreactions have been known between large-scale and lab-scale reactors: i) the biomass yield is reduced at large-scale; ii) more metabolic by-products are produced at large-scale; and iii) limiting substrate gradients are present at large-scale as measured at different heights in the bioreactor (Bylund et al., 1998; Xu et al., 1999). All these phenomena are classified under the term ‘scale-up effect’. Although scale-up is still regarded more as an art than a science (Humphrey, 1998), transport limitation is considered as one of the major factors responsible for phenomena observed at large-scale. For this reason, it is necessary to investigate both the biological and technological aspects of the system in the large scale. Prevention of slowing down deteriorative processes is required after liquid fertilizer was produced by aerobic biodegradation in order to maintain its quality during the period of circulation in market. Generally, the lower the pH, the less the chance that microbes will grow and cause spoilage. It has been known that organic acids can lower the pH and have a bacteriostatic effect (Zhuang et al., 1996). A number of other methods have also reported for microbial control (Agarwal et al., 1986; Stratham and Bremner, 1989; Curran et al., 1990). However, information of the preservation of liquid-fertilizer has been lacking so far. A means of a more long-term preservation of the liquid fertilizer is required. In this study, biodegradation of FMW was attempted in a large scale and suitability of biodegraded product was determined as a fertilizer. For this purpose, microorganisms were newly isolated and their characteristics of biodegradation were investigated in lab scale. Based on the lab-scale data, biodegradation of FMW was then carried out in a 1-ton

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bioreactor. To examine the fertilizing value of the biodegraded end product, analysis of amino-acid composition and tests of seed germination and root length were accomplished. The long-term preservation of the end-product was also studied.

EXPERIMENTAL Isolation of Useful Microorganisms The potential aerobically-degrading bacteria were isolated from commercial good-quality humus and from compost and leachate collected at three different sites of composting plants. The soil and compost samples (0.5 g each), and 0.5 ml of raw leachate sample were added into 5 ml of sterile 0.2% NaCl and agitated to obtain homogeneous suspension. One ml of each suspended liquid was pipetted into various 10 ml- tubes that contained 0.8% nutrient broth (pH 6.8), yeast-maltose medium (3 g·l-1 of yeast extract, 3 g·l-1 of malt extract, 5 g·l-1 of peptone, 10 g·l-1 of glucose, and 0.05 g·l-1 of ampicillin, pH 6.2) and Bennet’s medium (1 g·l-1 of yeast extract, 1 g·l-1 of beef extract, and 10 g·l-1 of glucose, pH 7.2). After one day incubation at both of 45℃, the liquid culture was spread onto the agar plate of each medium. The separated colonies formed on the plates were serially picked up and inoculated onto the fresh agar plates repeatedly until a pure isolate was obtained. All isolates were spread on four different agar plates: 1% skim milk agar for detection of proteolytic microorganisms; 3.215% spirit blue agar for detection of lipolytic microorganisms; and starch hydrolysis agar (5 g·l-1 of beef extract, 20 g·l-1 of soluble starch, 10 g·l-1 of tryptose, 5 g·l-1 of NaCl, and 15 g·l-1 of agar, pH 7.4) and cellulose agar (10 g·l-1 of cellulose powder, 1 g·l-1 of yeast extract, 0.1 g·l-1 of NaCl, 2.5 g·l-1 of (NH4) 2SO4, 0.25 g·l-1 of K2HPO4, 0.125 g·l-1 of MgSO4·7H2O, 0.0025 g·l-1 of FeSO4·7H2O, 0.025 g·l-1 of MnSO4·4H2O, and 15 g·l-1 of agar, pH 7.2) for detection of carbohydrate-degrading microorganisms, respectively. All agar plates were incubated under same conditions until change of color or a clear zone around each colony appeared.

Tests of Antagonism and Salt Effect on Growth Screening of potential bacterial antagonists against other isolated microorganisms was carried out by the use of perpendicular streak technique as described by Alippi and Reynaldi (2006). Each plate was incubated at both of 45°C for three days to allow the production of antagonistic substances and then checked for any growth inhibition of each isolate. After the antagonism test, the effect of salt concentration on cellular growth was investigated. Each screened cell was spread on four different agar (skim milk agar, spirit blue agar, starch hydrolysis agar and cellulose agar) plates containing various concentrations of 1, 2 and 3.5% NaCl additionally. All agar plates were incubated under the same conditions, and the effect of salt on the growth of each cell was verified by measuring difference in change of color or size of a clear zone around each colony.

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Identification of Useful Microorganisms Chromosomal DNA of the isolate was extracted from cells grown in the given medium with AccuPrep® Genomic DNA extraction kit (Bioneer), according to the manufacturer's instructions. PCR amplification of the DNA using the 27F (5'AGAGTTTGATCCTGGCTCAG-3') and 1492R (5'-GGTTACCTTGTTACGACTT-3') were performed with a PCR thermal cycler DICE model TP600 (Takara) as described by Kim et al. (2007), and the 16S rDNA genes were determined by the Macrogen Company (Seoul, Korea). Related sequences were searched against GenBank (National Center for Biotechnology Information, USA) using the Advanced BLAST similarity search option (Altschul et al., 1997) accessible from the homepage, http://www.ncbi.nlm.nih.gov/. BioEdit Sequence Alignment Editor version 5.0.9 was used to check alignment and remove all positions with gaps before calculating distances with DNAdist programme in PHYLIP (version 3.5c).

Lab-scale Aerobic Biodegradation Using the screened microorganisms, aerobic biodegradation was carried out in a 100 mlsyringe that served as the reaction vessel (Cho et al., 2006). Under supply of sterile oxygen, 0.2 g (wet weight basis) of mixed isolates (5% inoculums) were suspended in the syringe with 40 ml of the original FMW (pH 6.5±0.2) obtained from a fish-meal factory. To examine the possibility of combined wastewaters as a substrate for the production of liquid-fertilizer, milk wastewater (MW) generated from milk-processing factory or wasted broth (WB) generated after the cultivation of photosynthetic bacteria for biomass production was used with the original FMW at the dilutions of 10-fold and 32-fold. The effect of combined microorganisms (isolated microorganisms with a photosynthetic bacterium (PSB), Rhodobacter capsulatus) on biodegradation was also investigated. For faster biodegradation, the inoculated cells were previously acclimated for two days in the original FMW under an aerobic condition. The syringes prepared in this way were incubated in a shaking incubator at 45℃ and 180 rpm. The gas produced by the mixed microorganisms during incubation was analyzed by gas chromatography (GC). At the same time liquid broth was taken from the syringe to measure the concentrations of chemical oxygen demand- dichromate (CODCr) and total nitrogen (TN).

Large-scale Aerobic Biodegradation Large-scale biodegradation using original FMW was executed twice in a 1-ton bioreactor. The characteristics of the FMW are given in Table 1. The pH of FMW was not adjusted because it was always measured to be 7.0+0.3, and 30 l of isolated mixed microorganisms were seeded into the bioreactor filled with 600 l of original FMW. The bioreactor was operated at 42+4℃, and air from the blower was supplied into the bioreactor through ceramic disk-typed diffuser. Two blowers and 12 disks were used to supply adequate oxygen into the bioreactor. Ten-fold diluted ‘Antifoam 204’ was used when severe foams occurred during biodegradation. Samples from the bioreactor were collected periodically, and the

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concentration of dissolved oxygen (DO), oxidation-reduction potential (ORP) and pH were measured. Table 1. Characteristics of the original FMW Constituents

Concentration (mg·l-1)

CODCr

115,000±13,000

TN BOD5

15,400±1,300 68,900±7,600

NH4+-N

2,800±600

NO3--N

0

NO2--N

0

Seed Germination Test To evaluate the phytotoxicity of biodegradated FMW, seed germination test was carried out according to the method of Wong et al. (2001). Five milliliters of sample were pipetted into a sterile petri dish lined with Whatman #1 filter paper. Ten cress (Lepidium sativum) seeds were evenly placed in each dish. The plates were incubated at 25 ºC in the dark at 75% of humidity. Distilled water was used as a control. Seed germination and root length in each plate were measured at 72 h. The percentages of relative seed germination (RSG), relative root growth (RRG) and germination index (GI) after expose to wastewater treated were calculated as the following formula (Zucconi et al., 1981; Hoekstra et al., 2002):

RSG (%) = RRG (%) = GI (%) =

Number of seeds germinated in biodegraded wastewater Number of seeds germinated in control Mean root length in biodegraded wastewater Mean root length in control RSG × RRG

× 100

× 100

100

Preservation of Final Broth Various concentrations (0.5, 1 and 3%) of lactic acid were used to preserve four different types of biodegradated wastewaters. The preservation was carried out for six months in a 1 l plastic bottle, which was stored at room temperature. During the preservation, samples were periodically taken under a clean bench for the odor evaluation by panel. The panel was composed of twenty persons. The change of level of amino acids was also analyzed.

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Analyses The concentrations of cations (NH4+) and anions (NO2- and NO3-) were estimated by IC (Metrohm 792 Basic IC). The columns used in these analyses were Metrosep C2-150 and Metrosep Supp 5-150 for cation and anion, respectively. The concentrations of CODCr and TN concentrations were analyzed by the Water-quality Analyzer. The five days biological oxygen demand (BOD5) was analyzed by the OxiDirect BOD-System. The composition of amino acids in samples was analyzed at Feeds and Foods Nutrition Research Center in Pukong National University by our request. For determination of nitrogen and carbon dioxide gases, 20 μl samples (injection volume) were taken for GC/TCD (Perkin Elmer Instruments) analysis. The columns used were a 'molecular sieve 13X' and 'carboxen 1,000' for nitrogen and carbon dioxide, respectively. In analyses of both gases, the following conditions were equally applied: the carrier gas was helium at a flow rate of 30 ml·min-1 and the injector and the detector temperatures were 100 and 200℃, respectively. However, the oven temperature for nitrogen gas was 40℃, and that for carbon dioxide gas was 40℃ for 3 min initially then increased to 170℃ with a rate of 30℃·min-1.

RESULTS AND DISCUSSION Useful Microorganisms for Biodegradation of FMW Screening of useful microorganisms In our previous study (Kim et al., 2007), we finally developed seven isolates, which were able to produce no antagonistic substances against other microorganisms among forty-six isolates. The seven isolates were given the names as JB1 to JB7, respectively. All the cells were very motile in vegetative state and Gram-positive rods measuring 0.5-0.7 μm in width and 3-5 μm in length. In the experiment for examination of the salt effect on cellular growth, there was no effect on cellular growth of each microorganism at the concentrations of 1 and 2% of NaCl. However, the effect was distinct at 3.5%. The concentration of salt in a raw material (fish wastes) used for the production of fish meal varies, and thus the salt concentration of the original FMW can also vary. The salt concentration of the original FMW used in this study was measured to be much less than 1% (0.6±0.1%). It is concluded that the low salt concentration of FMW cannot affect the growth of the seven isolates.

Identification of the screened isolates Species-specific identification for the seven isolates could be derived using 16S-rDNA sequence analysis, since each species possesses one or more unique 16S-rDNA nucleotide regions. Approximately 1,500 bp sized- fragment of the 16S-rRNA gene of each isolate was amplified and sequenced. Each fragment band was confirmed by electrophoresis after performance of PCR. Sequence analysis of the 16S-rDNA gene and BLAST sequence comparison confirmed that the isolated strain JB1 was Bacillus subtilis (GenBank Accession

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Number: DQ219358), JB2 was Bacillus licheniformis (GenBank Accession Number: AY468373), JB3 was Brevibacillus agri (GenBank Accession Number: AY319301), JB4 was Bacillus coagulans (GenBank Accession Number: AF466695), JB5 was Bacillus circulans (GenBank Accession Number: Y13064), JB6 was Bacillus anthracis (GenBank Accession Number: AY138279) and JB7 was Bacillus fusiformis (GenBank Accession Number: AY548950) with similarity of 98-100%.

Metabolic characteristics of the isolates Metabolic characteristic of each isolate was characterized in a 100 ml syringe reactor, and is tabulated in Table 2. Among the seven isolates, the isolate JB7 had the highest values of maximum O2 consumption rate and maximum N2 production rate under an aerobic condition, which resulted in increase of pH during biodegradation. The isolate JB2 showed the similar characteristic with high maximum CO2 production rate. However, JB3, JB4 and JB5 had relatively lower O2 consumption rate with lower N2 and CO2 production rates. Interaction always occurs between diverse microbial populations, and both of them may benefit from the interactions. Coexistence in the mixed culture has been reported to occur only if between species competition is weaker than within species. Recently, a new model with introduction of mutualism between competitive species has been proposed (Zhang, 2003). Thus, mutualism among isolated microorganisms could promote coexistence and enhance the carrying capacity of the system, since they did not show any antagonism. Table 2. Metabolic characteristics of isolated microorganisms Isolated microorganism

Max. O2 consumption rate (mole·h-1)

Max. N2 production rate (mole·h-1)

Max. CO2 production rate (mole·h-1)

JB1

2.77

1.84

0.25

JB2

4.27

2.50

0.83

JB3

1.45

1.00

0.21

JB4

1.80

0.22

0.20

JB5

1.44

0.76

0.23

JB6

2.39

0.10

0.18

JB7

4.38

3.06

0.62

Lab-Scale Biodegradation Effects of oxygen and dilution ratio on biodegradation of FMW The experiment was carried out in a 100 ml syringe. In our previous study (Kim et al., 2007), it was found that the concentrations of CODCr and TN in original FMW were much more reduced by the seven isolates under the condition of O2 supplement, compared to those

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under the condition of no O2 supplement. With supplement of O2, the production of CO2 gas was increased with increase of N2, but only small bubble was produced in the syringe vessel without supplement of O2. The result indicates that the greater mineralization of the organic matter occurred under an aerobic condition. It is known that oxygen consumption is a general index of microbial metabolism (Tomati et al., 1996). Thus, the seven isolates showed more active mineralization of the organic matter under an aerobic condition. The different content of organic matter in FMW can affect the biodegradation, since cellular metabolism is dependent on substrate concentration (Maria et al., 2000). For this reason, the effect of dilution ratio of FMW on biodegradation was also examined in our previous study using a 100 ml syringe (Kim et al., 2007). The oxygen consumption rate by the seven isolates in the syringe vessel tended to increase when more diluted FMW was used as substrate, i.e., faster biodegradation occurred with more diluted FMW, which resulted in faster removal rates of CODCr and TN. The maximum rates of gas productions of CO2 and N2 during biodegradation were the highest with 8-fold diluted FMW, and the microbial population also tended to increase with more diluted FMW.

Biodegradation of FMW in a lab-scale bioreactor In our previous study (Kim et al., 2007), we examine the characteristics of aerobic biodegradation of FMW in a 5l- bioreactor starting on 8-fold diluted FMW. The removal percentages of CODCr and TN were more prominent with slight decrease of CODCr/TN ratio, compared with the results of the syringe experiment using the same diluted FMW. This is because oxygen is able to be supplied more sufficiently into a bioreactor, but not in the syringe vessel. It has been known that the COD/N ratio may influence biomass activity, and therefore on the metabolic pathways of organic matter utilization (Ruiz et al., 2006). Based on this information, the cell activity and metabolic pathways of the seven isolates may be maintained steadily during the period of this experiment. The change of pH was not considerable and the weight of sludge decreased approximately half, which means that some fractions of suspended solids were biodegraded. Reduced phytotoxicity of final FMW broth was attained from aerobic biodegradation in a 5l- bioreactor when more diluted (32-fold) FMW was used. The trend in reduction of CODCr or TN was not much different, and pH was also not changed considerably. However, a noticeable result in this experiment was that a strong unpleasant smell (mainly a fishy smell) mostly disappeared in the end. Phytotoxicity of final broth Sufficient aeration promotes the conversion of organic matters into nonobjectionable, stable end products such as CO2, SO42-, NO3-, etc. An incomplete aeration may result in accumulation of organic acid, thus giving trouble to plant growth if the fertilizer is incorporated into the soil (Jakobsen, 1995). To examine the fertilizing value of the final broth taken from a 5-l bioreactor, phytotoxicity assays were accomplished in our previous study (Kim et al., 2007). The effect of the final broth on the germination was found to be not pronounced, since all cress seeds used in the experiment were germinated in one day. However, it has been reported that seed germination is regarded as a less sensitive method than root elongation when used as a bioassay for the evaluation of phytotoxicity (Wang and Kentri, 1990). Instead, the germination index (GI), which combines the measure of relative seed germination (RSG) and relative root growth (RRG) of cress seed (Lepidium sativum),

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has been reported to be the most sensitive parameter used to evaluate the toxicity (Zucconi et al., 1981). When phytotoxicity of the final broth of original FMW was assayed, its average GI was only 8.0%. The reduction in GI indicates that some characteristics existed had an adverse effect on root growth. This may be attributed to the release of high concentrations of ammonia and low molecular weight organic acids (Wong, 1985; Fang and Wong, 1999), since cress used in this study is known to be sensitive to the toxic effect of these compounds (Fuentes et al., 2004). The values of GI (%) tended to increase with increase of dilution ratio of the final broth. At the dilution ratio of 32, the average value of GI was found to be over 50%. A GI of 50% has been used as an indication of phytotoxin-free compost (Zucconi et al., 1985). According to this GI criterion, the final broth of biodegradation using 8-fold diluted FMW required more mineralization to reach stabilization. Phytotoxicity caused by organic compounds can be remedied by increasing the period of aerobic decomposition (Wong et al., 2001). Instead of increasing the period of biodegradation, the concentration of substrate had to be reduced to remedy the phytotoxicity, since cellular metabolism is dependent on substrate concentration (Maria et al., 2000). Consequently, more diluted FMW was required for the further stabilization of the organic matter to maintain the long-term fertility in soil. Phytotoxicity was also assayed on the final FMW broths at various dilution ratios in our previous study (Kim et al., 2007). The GI value of the final broth biodegraded on 32-fold diluted FMW was higher than that of the final broth biodegraded on 8-fold diluted FMW at the same dilution ratio. This indicates that the degree of mineralization was higher in the biodegradation using 32-fold diluted FMW and this resulted in high value of GI, i.e., GI tended to increase as the content of mineralized organic matter in FMW increased by biodegradation (Fuentes et al., 2004). The final broth of biodegradation using 32-fold diluted FMW was found to require only two-fold dilution to reach stabilization.

Composition of amino acids in final broth Amino acids are an essential part of the active fraction of organic matter in a fertilizer. The growth of plants depends ultimately upon the availability of a suitable balance of amino acids, and their composition might also be used as a means of assessing biodegradation. From this point of view, the amino-acid composition of the biodegraded FMW taken from a 5-l bioreactor was analyzed. In our previous study (Kim et al., 2007), the results clearly showed that the amino-acid composition in the final broth (12.54 g·100g sample-1) was almost twice that of non-biodegraded FMW. The higher content of amino acids in the final broth is probably due to the higher degree of mineralization of FMW, which indicates release of more nutrients available for plants. The amino-acid composition in the final broth was also higher in comparison with that of a commercial fertilizer for horticultural plants. Moreover, the content of sulfur-containing amino acids, cysteine and methionine, were much higher. It has been reported that the sulfur-containing amino acid, methionine is a nutritionally important essential amino acid and is the precursor of several metabolites that regulate plant growth (Amir et al., 2002). Consequently, the levels of amino acids in the final broth are comparable to those in a commercial fertilizer.

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Biodegradation on combined wastewater The dairy industry generates a large amount of effluent ranging from 0.2 to 10 L per L of processed milk. Wastewaters from the dairy industry are polluting chiefly because of the organic matter they contain and they should be treated before discharge into surface waters. Conventional treatment of dairy wastewater involves aerobic processes, since fats, lactose and proteins are all easily degraded by bacterial populations (Samkutty et al., 1996). For industrial reutilization of milk waste (MW), biodegradation on the combined wastewater of FMW and MW was investigated in a 100-ml syringe. In the experiment 10-fold and 32-fold diluted combined wastewaters were used, and the results are shown in Fig. 1 and Fig. 2. As the

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mixing ratio of FMW to MW decreased, the organic strength of the combined wastewater was weaker. In both experiments, the increase of the amount of MW in the combined wastewater did not give any remarkable effect on the production of N2 or CO2. Removal of CODCr and TN was rather reduced, compared to that of control. This indicates that isolated microorganisms seem to be not fit for biodegradation of MW, probably due to its different organic component. To reutilize FMW with MW together, further study is required to develop some useful microorganisms, especially for biodegradation of MW.

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Biodegradation by combined microorganisms It has been reported that photosynthetic bacteria (PSB) can consume various types of organic substrates, nitrogenous and phosphorous compounds simultaneously, with a relatively high growth rate (Sasaki et al., 1998). For this reason, the effect of the combined microorganisms (by the addition of PSB to isolated microorganisms) on biodegradation was investigated in a 100-ml syringe, and the result is shown in Fig. 3. To maintain high activity of PSB during biodegradation, the experiment was carried out on the combined wastewater (1:1 mix of FMW and wasted broth (WB) generated after the cultivation of PSB), since PSB has been reported to excrete growth-related metabolites to the abiotic phase (Kim et al.,

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2001). The increase of the amount of PSB in the combined microorganisms resulted in reduced production of N2 and CO2 and reduced removal of CODCr and TN as well, compared to that of control. It has been known that cell interaction can be very complex in mixed culture (Purtschert and Gujer, 1999). The best use of PSB was tried, but the advantage of PSB addition was not shown in this case.

LARGE-SCALE BIODEGRADATION Biodegradation of FMW in a large-scale bioreactor The large-scale FMW biodegradation from its lab-scale were attempted in a 1-ton bioreactor twice, and the results are shown in Fig. 4. In first trial (Fig. 4A), DO level in the bioreactor was maintained low (less than 1 mg·l-1), and pH increased up to 8.5. The ORP during biodegradation showed negative values with steady decrease from the beginning of the experiment. This implies that the supply of oxygen cannot meet the demand of oxygen by isolated microorganisms as they grew. Generally, it has known that DO level in a bioreactor should be maintained over 1 mg·l-1 for aerobic fermentation (Tohyama et al., 2000). The problem is to estimate the proper aeration rate in the large vessel.

To increase DO concentration level in the bioreactor, additional blower and oxygen diffuser were installed. Thus, the aeration rate was increased from 320 to 1,280 l·min-1. The result of the second trial is shown in Fig. 4B. Under the improved condition of aeration, DO level in the bioreactor could be maintained over 2 mg·l-1 during two days biodegradation, but thereafter it decreased. The pH was maintained in a range of 6.2-7.0, and the values of ORP were mostly positive during biodegradation. This trend showed different from that of the first

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trial, and the odor disappeared greatly. From these results, maintenance of DO level found to be very important and ORP could be a key parameter to operate biodegradation of FMW in a plant-scale. ORP was reported to be used as a controlling parameter for regulation of sulfide oxidation (Khanal and Huang, 2003), and on-line monitoring of ORP has been proved to be a practical and useful technique for process control of wastewater treatment systems (Yu et al., 1997). In aerobic processes, oxygen is a key substrate and because of its low solubility in aqueous solutions a continuous transfer of oxygen from the gas phase to the liquid phase is decisive for maintaining the oxidative metabolism of the cells. Thus, process optimization is required in a large-scale operation, especially aeration rate in this case. The reduction of CODCr or TN in the both trials was not much different.

Final broth as liquid fertilizer The success of the scale-up process for the production of liquid fertilizer from FMW can be verified by determining the composition of amino acids. The composition of amino acids between the final broths taken from first and second trials was analyzed and compared, and its result is tabulated in Table 3. The amino-acid composition of the second trial was higher than that of the first trial. This implies that better biodegradation under supply of sufficient oxygen resulted in increase of content of amino acids. Thus, steady supply of sufficient oxygen into large-scale bioreactor is very important. The level of total amino acids was lower, compared to that in commercial liquefied fertilizer. Especially the levels of apartic acid, glutamic acid, alanine and lysine were short, whereas those of proline and glycine were higher. The difference may be due to the limitation of oxygen, since the DO level could not be steadily maintained. Or the composition of the original FMW may be much different, since it was found to depend on the nature of the raw material processed in the factory (Kim et al., 2007). Phytotoxicity assays were accomplished for liquid-broths taken at different times of biodegradation during the second trial. As shown in Fig. 5, the values of GI tended to increase with increase of dilution ratio of the liquid broths. The value of GI also tended to increase with more biodegraded broth as cultivation time elapsed. It has reported that GI tends to increase as the content of mineralized organic matter increases (Fuentes et al., 2004). Organic materials hold great promise due to their local availability as a source of multiple nutrients and ability to improve soil characteristics. Thus, it implies that more degraded organic compounds were present as cultivation time elapsed. In the case of experiment with liquid broth degraded for 3 days, only the half of cress seeds was germinated when 10-fold diluted broth was used, and its GI value was less than 10% with low root elongation. More diluted (100-fold) liquid broth resulted in increase of GI, and GI reached over 80%. This is enough to reach stabilization of the organic matter to maintain the long-term fertility in soil, according to the GI criterion (Zucconi et al., 1985). In case of commercial liquid-fertilizer, more than 100 dilutions are general for use in soil. A field experiment concerning the effect of final broth on soil fertility will be proceeding after this work.

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Table 3. The comparison of amino-acid composition between final broths of first and second trialsa Source of final broth Amino acid First trial Second trial Aspartic acid 0.32 0.49 Threonine 0.18 0.17 Serine 0.19 0.21 Glutamic acid 0.54 0.78 Proline 0.50 0.50 Glycine 0.47 1.06 Alanine 0.32 0.60 Valine 0.15 0.10 Isoleucine 0.10 0.14 Leucine 0.20 0.24 Tyrosine 0.07 0.06 Phenylalanine 0.19 0.09 Histidine 0.09 0.20 Lysine 0.29 0.24 Arginine 0.31 0.24 Cystine 0.02 0.04 Metionine 0.13 0.06 Total 4.07 5.22 a

composition of amino acids was based on fry weight (g•100g sample-1)

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Preservation of Final Broth Effect of lactate on preservation The effect of lactate addition on the quality of final broths was investigated, and the results are tabulated in Table 4. Four different types of biodegraded wastewaters were stored by the various concentrations of lactate as a reagent for preservation. At the beginning of the experiment, FMW-1 emitted ammonia smell, whereas FMW-2 emitted relatively light ammonia smell. The smell seemed to depend on the quality of the biodegraded wastewater. In cases of control and addition of 0.5% lactate, unacceptable odors were produced from all types of wastewaters within 45 days by putrefaction. When 3% lactate was added to each biodegraded wastewater, the soy sauce-like smell could be emitted and retained for six months whether the quality of the biodegraded wastewater was good or not. The addition of 1% lactate could preserve the biodegraded FMW-2 for six months with the soy sauce-like smell. It has been reported that sweet-smell soy sauce had eighteen different free amino acids (Jingtian et al., 1988). This indicates that good-quality FMW-2 could be preserved for six months without any putrefaction. However, FMW-1, which was degraded poorly due to deficiency of oxygen, produced strong odor after two months. Addition of 1% lactate caused the change of pH from 7.8 to 5.7, with increase of ORP from - 10 to 73.2 mV. Consequently, lactate prevented the growth of spoilage microorganisms keeping their microbial counts steady and pH values within the acid region. Table 4. Effect of lactate addition on the quality of biodegraded final-broths during their preservation at room temperature Type of wastewatera

FMW-1

FMW-2

COMB-1

COMB-2 a

Addition of lactate (%) 0 (control) 0.5 1 3 0 (control) 0.5 1 3 0 (control) 0.5 1 3 0 (control) 0.5 1 3

0 Ab A A A L L L L A A A A A A A A

10 A A A Lb A L S S A A A L A A A L

20 A A A Sb A A S S A A A S A A A S

30 Ob O A S O A S S O O A S O O A S

Storage time (day) 45 60 75 90

120

150

180

A S

O S

S

S

S

S

S

O S S

S S

S S

S S

S S

S S

S S

A S

O S

S

S

S

S

S

A S

O S

S

S

S

S

S

biodegraded wastewater: ‘FMW-1’ was degraded under deficiency of oxygen (Fig. 4A); ‘FMW-2’ was degraded under relatively adequate supply of oxygen (Fig. 4B); COMB-1 was degraded on the combined wastewaters of 90% FMW-2 and 10% MW; and COMB-2 was degraded on the combined wastewaters of 90% FMW-2 and 10% MW containing EM (effective microorganisms). b symbols represent different smell: ‘A’ means ammonia smell; ‘O’ means odor by putrefaction; ‘L’ means relatively light ammonia smell; and ‘S’ means soy sauce-like smell.

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The effect of lactate addition on preservation of the combined wastewaters was investigated. Use of effective microorganisms (EM) inoculum along with organic materials has been known to be an effective technique for stimulating supply and release of nutrients from these nutrient sources. Some studies have shown that the inoculation of agro-ecosystems with EM cultures can improve soil and crop quality (Daly and Stewart, 1999; Hussain et al., 1999; Khaliq et al., 2006). Thus, EM is known to be an additive for optimizing all other amendments and practices used for crop production. Our study showed that better preservation was not found on the combined wastewater, COM-1 or COM-2 at the same concentration of lactate added. This implies that the role of EM in preservation was very weak. In conclusion, 1% lactate is a good preserving reagent for good-quality liquid fertilizer.

Change of amino-acid composition The change of amino-acid composition was investigated along with the storage time, and the results are tabulated in Table 5. The levels of several amino acids slightly increased as storage time elapsed. Due to the production of amino acids, the pH of the liquid broth decreased further to 5.3 after 6 months. This implies that lactate could make the isolated microorganisms maintained at their minimum activity, which resulted in emission of soy sauce-like smell. Table 5. Change in amino-acid composition during preservation by addition of lactatea Storage time (day)

Amino acid 0

30

180

Aspartic acid

0.49

0.49

0.51

Threonine

0.17

0.18

0.21

Serine

0.21

0.21

0.21

Glutamic acid

0.78

0.79

0.84

Proline

0.50

0.51

0.52

Glycine

1.06

1.07

1.08

Alanine

0.60

0.61

0.62

Valine

0.10

0.11

0.14

Isoleucine

0.14

0.15

0.18

Leucine

0.24

0.25

0.28

Tyrosine

0.06

0.07

0.10

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Table 5. Change in amino-acid composition during preservation by addition of lactatea Storage time (day) Amino acid 0

30

180

Phenylalanine

0.09

0.10

0.15

Histidine

0.20

0.21

0.22

Lysine

0.24

0.24

0.28

Arginine

0.24

0.24

0.24

Cystine

0.04

0.04

0.06

Metionine

0.06

0.06

0.07

Total

5.22

5.33

5.71

a

composition of amino acids was based on fry weight (g•100g sample-1)

CONCLUSION To reutilize the wastewater generated during the process of fish-meal production (FMW), seven thermophilic microorganisms were newly isolated and their characteristics of aerobic biodegradation of FMW were examined in a lab-scale bioreactor. It clearly showed that the amino-acid composition (12.54 g·100g sample-1) in the final broth of the biodegradation using 8-fold diluted FMW was almost twice that of non-biodegraded FMW. The levels of amino acids in the final broth were also comparable to those in a commercial fertilizer. When more (32-fold) diluted FMW was used as a substrate, phytotoxicity of biodegraded final broth was further reduced with disappearance of a strong unpleasant smell in the end. The results of the lab-scale biodegradation suggest the promising potential of biodegraded FMW for the production of fertilizer. From the laboratory results, a large-scale biodegradation using the original FMW was designed and carried out in a 1-ton reactor. During the biodegradation, the concentration of DO in the liquid broth was closely related to pH and the oxidation reduction potential (ORP) as well. Keeping a low level of DO resulted in both the increase of pH and the decrease of ORP with strong smell of ammonia. Under the maintenance of DO level over 1 mg·l-1, the initial fishy smell from the FMW was converted to a pleasant smell in the end with production of fairly good content of amino acids. Therefore, the DO level in the liquid broth was found to be decisive influence on the quality of final fermented broth, and ORP to be a key operation parameter in biodegradation of FMW. The final broth taken from the bioreactor is required to be maintained its quality as a liquid fertilizer, during the period of circulation in market. The addition of 3% lactic acid could preserve the final broth well for six months, whereas the addition of lower concentrations of lactic acids could not preserve properly and resulted in putrefaction in the end. When a good-quality final broth was used, the addition of

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1% lactic acid could preserve it for six months with amino-acids enriched soy sauce-like smell. From the results of the large-scale biodegradation, the reutilization of FMW is expected to yield high economic value.

ACKNOWLEDGEMENTS This research was supported by a grant (B-2004-12) from Marine Bioprocess Research Center of the Marine Bio21 Center funded by the Ministry of Maritime Affairs and Fisheries, Republic of Korea.

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Kim J. K, Kim J.-B., Cho, K.-S. and Hong, Y.-K., 2007. Isolation and identification of microorganisms and their aerobic biodegradation of fish-meal wastewater for liquidfertilization. International Biodeterioration Biodegradation 59, 156-165. Kim, J.-K., Yun, J.-S. and Ryu, H.-W., 2001. Simultaneous determination of δaminolevulinic acid, porphobilinogen, levulinic acid and glycine in culture broth by capillary electrophoresis. Journal of Chromatography A, 938, 137-143. Maria, G., Maria, C., Salcedo, R. and Azevedo, S., 2000. Databank transfer-of-information, shortcut and exact estimators used in the wastewater biological treatment process identification. Computers & Chemical Engineering 24, 1713-1718. Martin, A. M., 1999. A low-energy process for the conversion of fisheries waste biomass. Renewable Energy 16, 1102-1105. Park, E., Enander, R., Barnett, S. M. and Lee, C., 2001. Pollution prevention and biochemical oxygen demand reduction in a squid processing facility. Journal of Cleaner Production 9, 341-349. Purtschert, I. and Gujer, W., 1999. Population dynamics by methanol addition in denitrifying wastewater treatment plants. Water Science and Technology 39, 43-50. Ruiz, G., Jeison, D. and Chamy, R., 2006. Development of denitrifying and methanogenic activities in USB reactors for the treatment of wastewater: Effect of COD/N ratio. Process Biochemistry 41, 1338-1342. Samkutty, P. J., Gough, R. H. and McGrew, P., 1996. Biological treatment of dairy plant wastewater, J. Environ. Sci. Health A31, 2143–2153. Sasaki, K., Tanaka, T. and Nagai, S., 1998. Use of Photosynthetic Bacteria for the Production of SCP and Chemicals from Organic Wastes. Bioconversion of Waste Materials to Industrial Products, 2nd ed., Martin, A. M. (Ed), Blakie Academic & Professional (Chapman & Hall), London, New York, Tokyo, pp. 247-290. Stratham, J. A. and Bremner, H. A., 1989. Shelf-life extension of packed seafoods: A summary of research approach. Food Austria 41, 614-620. Tohyama, M., Takagi, S. and Shimizu, K., 2000. Effect of controlling lactate concentration and periodic change in DO concentration on fermentation characteristics of a mixed culture of Lactobacillus delbrueckii and Ralstonia eutropha for PHB production. Journal of Bioscience and Bioengineering 89, 323-328. Tomati, U., Galli, E., Fiorelli, E. and Pasetti, L., 1996. Fertilizers from composting of olivemill wastewaters. International Biodeterioration and Biodegradation 38, 155-162. Wang, W. and Kentri, P. H., 1990. Comparative seed germination tests using ten plants species for toxicity assessment of a metal engraving effluent sample. Water, Air and Soil Pollution 52, 369-376. Wong, M. H., 1985. Phytotoxicity of refuse compost during the process of maturation. Environmental Pollution (Series A) 40, 127-144. Wong, J. W. C., Mak, K. F., Chan, N. W., Lam, A., Fang, M., Zhou, L. X., Wu, Q. T., and Liao, X. D., 2001. Co-composting of soybean residues and leaves in Hong Kong. Bioresource Technology 76, 99-106. Xu, B., Jahic, M., Blomsten, G. and Enfors, S.-O., 1999. Glucose overflow metabolism and mixed acid fermentation in aerobic large-scale fed-batch processes with Escherichia coli. Applied Microbiology and Biotechnology 51, 564–571.

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Yu, R.-F., Liaw, S.-L., Chang, C.-N., Lu, H.-J. and Cheng, W.-Y., 1997. Monitoring and control using on-line orp on the continuous-flow activated sludge batch reactor system. Water Science and Technology 35, 57-66. Zhang, Z., 2003. Mutualism or cooperation among competitors promotes coexistence and competitive ability. Ecological Modelling 164, 271–282. Zhuang, R.-Y., Huang, Y.-W. and Beuchat, L. R., 1996. Quality changes during refrigerated storage of packed shrimp and catfish fillets treated with sodium acetate, sodium lactate or propyl gallate. Journal of Food Science 61, 241-244. Zucconi, F., Forte, M., Monaco, A. and Beritodi, M., 1981. Biological evaluation of compost maturity. Biocycle 22, 27-29. Zucconi, F., Monaco, A., Forte, M. and Beritodi, M., 1985. Phytotoxins during the stabilization of organic matter. In: Gasser, J. K. R. (Ed.), Composting of agricultural and other wastes. Elsevier, London, pp. 73-86.

In: Environmental Biodegradation Research Focus Editor: B. Y. Wang, pp. 239-257

ISBN: 978-1-60021-904-7 © 2007 Nova Science Publishers, Inc.

Chapter 9

METHODS IN STUDY OF BIODEGRADATION OF WATER INSOLUBLE POLYMER MATERIALS Marek Koutnya,∗ and Anne-Marie Delortb a

Tomas Bata University in Zlin, Faculty of Technology, Department of Environmental Protection Engineering, Czech Republic. b Laboratoire de Synthese Et Etude de Systemes a Interet Biologique (SEESIB), Ensemble Universitaire des Cezeaux, Universite Blaise Pascal, France

ABSTRACT Increasing waste disposal problems from polymer wrapping materials have resulted in constant endeavors to replace recalcitrant materials with biodegradable alternatives. The biodegradability of these materials can often simply be limited, or the processes involved are relatively slow, causing complications when applying standard methodologies and thus promoting the development of customized testing protocols. Moreover, some properties of these materials, especially their water insolubility, require further adaptations to conventional methods. This chapter brings together data found in literature, along with the personal findings and experiences of the authors. A broad variety of experimental methods has been used and described in literature, which are dependent on authors’ expertise or the availability of particular techniques in their laboratory. Here, a comprehensive overview of current and the most prominent techniques is provided. These include spectroscopic techniques to examine changes in the material; NMR, MS and separation methods for investigating compounds released from the material; electron microscopy methods; optical fluorescence microscopy that enables surfaces of the material and eventual microbial colonization to be visualized; different methods for biomass quantification and indicators of metabolic activity; and various ways of monitoring carbon dioxide production. The relevance of the methods for studying biodegradable synthetic polymer materials was analyzed, compared and then ∗

Send correspondence to Marek Koutny; Tel.: +420 576 031 221; Fax: +420 577 210 722; E-mail address: [email protected]

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Marek Koutny and Anne-Marie Delort critically evaluated. Original microphotography and original data have been introduced to illustrate the text.

1. INTRODUCTION In recent years, a growing interest in biodegradable polymeric materials has become apparent. Material scientists have created new compositions and blends of known biodegradable materials, and invented new principles and ideas, in order to make existing conventional materials biodegradable. Immediately afterwards, these materials are introduced on the market, often with only vague information on their real environmental fate. It is obvious that commercial subjects often disregard laborious and time consuming testing if there is no immediate pressure on them from legislation and government authorities and which is also true, if no clear and broadly accepted methodology can be perceived. The same ambivalence may be observed in scientific community, where opinion on whether a result from a particular method should be accepted as sufficient proof of biodegradability is now only slowly emerging. Even from recent years, papers on the topic can be found in respected peer-reviewed journals, where authors claim biodegradability by virtue of observations, which are doubtful at best. Research in the biodegradation of solid water insoluble polymeric materials must be considered as multidisciplinary. As such, it anticipates a need to adopt, use and develop expertise in a variety of techniques, which span from ones concerning the testing of mechanical properties, as well as those employed in the field of solid phase physics and analytical chemistry, to microbiology and biochemistry methods. In this area, there is no discernable absolute and most informative method, and any search in literature reveals a broad spectrum of diverse approaches utilized by different authors. A somewhat specific place in the field is reserved for procedures for monitoring CO2 production, in relation to the fact that CO2 is the final product of organic material mineralization. These methods are discussed in the last chapter. Considering the breadth of techniques employed, the only way of reaching correct conclusions is through open-minded and critical evaluation of obtained data. Since biodegradation processes for the material of interest are frequently very slow, the signals induced are weak and extreme care must be taken in their interpretation. The results obtained from different methods utilized during a study have to be compared, and their interpretations consistent, with a final conclusion on the material’s biodegradability. Roughly the methods can be divided into two categories. The first of these comprises methods intended to study changes in the material. The near complete range of techniques and characteristics available in polymer science was used to monitor biodegradation, including mechanical properties, various spectroscopic measurements, the degree of crystallinity assessed by different methods, changes in swelling, gas permeability, calorimetric techniques, and even rheological properties. In order to consider the usefulness of a particular method, it must be taken into account that biodegradation usually only takes place in a thin surface layer for insoluble samples. Therefore, taking mechanical properties as an example, e.g. frequently measured tensile strength, these may not document well biotic processes in such a case, being much more influenced by abiotic degradation progressed in

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the whole volume of sample. On the other hand, in many studies it was often very important to follow mechanical properties even when they are not directly related to biodegradation, because they are critical for evaluating the serviceability of the material. The results of the methods listed above can be also influenced by presence of biofilm on a sample’s surface, which is not easy to remove and which is not present on control samples. Carefully designed experiments with a reasoned set of control samples are necessary to distinguish abiotic degradation processes and biodegradation. In general, more attention must be paid while interpreting data and evaluating their significance when there is no clear and theoretically justified link between a physical property and a mechanism of biodegradation. The second group of methods includes those traditionally used by microbiologists. They are intended to monitor living biomass and processes carried out by microorganisms. Here, a perspective of the introduction of new and performing methodologies can be anticipated, since this is an area which takes advantage of rapid progress in life sciences. However, the classifications described above could not be followed in the structure of the chapter presented, as one individual method can frequently provide information both on the material and the interacting microorganisms. In particular, the scope of this chapter is determined by the authors’ experience gained during their studies of biodegradation of polyolephinic materials with prooxidant additives. Nevertheless, the facts presented can be broadly applied to other materials, especially those having character of water insoluble polymers or polymer composites.

2. PRELIMINARY OPERATIONS - SAMPLE PREPARATION Biodegradation of water insoluble materials takes place on the surface. Therefore, the form of samples is extremely important for the results of tests, particularly the ratio of the material surface to its volume. A film format- as thin as possible - appears to be the most suitable. With respect to the character of the material, the ideal thickness of film could be up to 100 μm. This is because it allows for the use of transmission methods like transmission FTIR or transmission optical microscopy in such instances. If the surface-volume ratio seems to be too unfavorable, the material must be disintegrated. The preferred way of doing this could be grinding it in a ball mill at the temperature of liquid nitrogen. Along with biodegradation, abiotic degradation processes also progress in the sample material (Arnaud et al., 1994). Abiotic degradation can contribute to overall degradation and can even eventually end up being the rate controlling step of the overall process. When this occurs, it can sometimes be necessary to carry out preliminary abiotic degradation of the sample to reach the level at which subsequent further biodegradation by microorganisms is possible. Consequently, advantage can be taken of the wealth of knowledge gathered on material aging, especially from artificial accelerated weathering methods. With help of such techniques, samples can be prepared with a level of abiotic degradation that corresponds to a predetermined period of aging in a selected natural environment under normal physical conditions. The sample, preprocessed in this way, can be then subjected to a biodegradation test (Wiles and Scott, 2006).

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Prior to biodegradation testing, a sample should be sterilized to ensure the biodegradation process is mediated by microorganisms that were voluntarily introduced into the experiment in a controlled way. However, there remains a problem that many methods of sterilization consist of a treatment including relatively drastic physical conditions, and/or with chemical agents which can unpredictably influence the physical or chemical nature of the sample, especially its level of abiotic degradation. Residuals of the disinfectant agent can also be paradoxically utilized by microorganisms as a carbon source and thus cause false positive results in biodegradation tests. It would appear that finding a suitable sterilization technique is a rather difficult task. In several studies, 95 % ethanol has been used as the relatively nonaggressive but insecure disinfecting agent. This offers the advantage that it can subsequently be evaporated out of the sample easily. For some analytical approaches intended to investigate changes on the material surface, it might prove indispensable to remove any biofilm covering the surface. Treatment with a suitable detergent might be useful for achieving this, such as the relatively mild non-ionic detergent Tween, or if such treatment is not sufficient, the stronger but rough sodium dodecyl sulfate (Orr et al., 2004) can be used. In addition, detergent action can be supported by ultrasound treatment.

3. IMAGING METHODS 3.1 Optical microscopy When the sample is in the form of a sufficiently thin film, transmission optical microscopy can be used to observe microorganisms on its surface. Staining is necessary in order to visualize cells and hyphae. Suitable techniques for this include Gram stain and, even more simply, with saphranine or basic fuchsin dyes (Koutny et al., 2006). Direct observation without staining, e.g. in phase contrast, could be difficult because fine unhomogeneities as well as eventual particles of filler are highlighted, in addition to microbial cells, which are then hard to distinguish. Prior to staining, it is best to fix microorganisms with a conventional fixative solution, for example those containing formaldehyde and/or glyceraldehyde as active agents. Optical transmission microscopy can also, to some extent, bring some additional information on cracks and other defects in the material, as well as indications of its deterioration and/or biodeterioration. The main benefit of optical microscopy appears to particularly lie in the relative rapidity of the technique in contrast to laborious and time consuming sample preparation methods for electron microscopy. First and foremost, it can be used as a simple and rapid control method during long incubations of samples.

3.2 Epifluorescence microscopy The sample can, of course, also consist of some thicker and/or non-transparent material. If so, microorganisms on the surface can be observed by epifluorescence microscopy following staining with suitable fluorescent dye. One example is acridine orange, which was

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successfully used for bacterial biofilm visualization on the surface of polyethylene film (Bonhomme et al., 2003). Obtained image data were subsequently transformed via a computer program and expressed in percentages of surface coverage. In this way, the time dynamic of biofilm propagation on the material’s surface could be evaluated. In the case described this method proved to be very effective. However, for another sample of oxidized polyethylene, it was shown that acridine orange and another conventionally used fluorescent dye DAPI (4'-6-diamidine-2-phenyl indole) manifested very strong non-specific binding to the sample material. Additionally, even the unstained sample material exhibited relatively strong fluorescence. Therefore, some limitations of the technique do exist. Epifluorescence microscopy can also help obtain information about biofilm viability (Sivan et al, 2006). The authors of that particular paper used the LIVE/DEAD BacLight Bacterial Viability Kit (Molecular Probes) to differentiate between live and dead cells in biofilm, which can be observed as cells of different color after applying the staining protocol according to the kit manufacturer’s instructions. When working with fluorescent dyes, strict safety precautions must be taken because the substances are often classified as carcinogenic.

Fig. 1. Rhodococcus rhodochrous on the surface of oxidized polyethylene film. Bottom images show detailed view of the upper ones.

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3.3 Scanning electron microscopy Scanning electron microscopy makes it possible to obtain sufficiently contrasted threedimensional images of microorganisms on a sample’s surface (Fig. 1). It enables examination of surface coverage with biofilm as well as morphology of microorganisms and their colonies. The enhanced capacity of microorganisms to attack the material can be exhibited by the penetration of hyphae into the material and colonization of cracks. Also, important data may be derived from observing the material itself, e.g. various changes on its surface, as well as the appearance of plaque and cracks. Prior to observing surface morphology, the biofilm should be removed. However, this can prove difficult to achieve in many cases with an acceptable non-aggressive treatment, as mentioned above. Nonetheless, the sample must be always compared with the identically treated abiotic blank. Sample preparation for electron microscopy techniques is a rather complicated procedure comprising sample fixation, dehydratation, and contrast enhancement (incorporating higher atomic weight elements). During the process, various artifacts can arise on the surface, such as various structures formed from the crystals of employed chemicals (Fig. 2).

Fig. 2. Possible artifacts of sample (oxidized PE film) preparation for scanning electron microscopy. a,b, samples after incubation without presence of microorganisms, still some shapes can evoke presence of bacterial cells or filaments. c, filaments of Nocardia asteroides covered by unknown substance probably during sample preparation process. d, hyphae of Cladosporium cladosporoides disintegrated probably during sample preparation.

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4. SPECTRAL METHODS 4.1 Fourier transformed infrared spectroscopy Infrared spectroscopy can be very informative and is frequently used in the majority of studies on polymer materials. It is capable of sensitively identifying changes in a material to the level of individual chemical bonds, and these may even be quantatively evaluated. Changes in infrared spectra recorded during the course of incubations can be used to monitor shifts in the relative abundance of selected chemical bonds, as well as the formation of new chemical bonds and functional groups. The method is considered one of the most prominent in monitoring abiotic degradation of polymeric materials. Applications of the methods to investigate biotic and/or abiotic changes in different polymeric materials is a discrete and rather broad topic, so cannot be discussed here in great detail. As regards polyethylene and other polymers, oxidative degradation appears to be the most important, characterized by the incorporation of oxygen into polymer molecules. The process can be observed by an increase in the signal of carbonyl groups at about 1712 cm-1. The growth of the peak has previously been linked with the formation of abiotic oxidation products, leading to the oxidative disruption of polymeric chains, where the new ends are terminated with carboxylic acid groups (Fig. 3) (Arnaud et al., 1994).

Fig. 3. Growth of carbonyl peak in transmission FTIR spectra during abiotic degradation of polyethylene film with prooxidants at 70°C. a, 0 days; b, 25 days; 67, days.

Another important functional group, probably formed during polyethylene oxidation, are esters. Their signals might be found close to 1740 cm-1 and explained away by the stretching vibration of the CO bond (Jakubowicz, 2006).

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When using ordinary transmission FTIR, one must take into account an important limitation. For water insoluble polymers, biodegradation normally takes place on the surface of the material. Thus, important chemical changes may be expected in a surface layer only a few micrometers thick, leading to rather poor sensitivity of transmission FTIR due to a very short optical path of infrared beam in a biodegraded layer. Therefore, the method is more suitable to monitor abiotic changes, where it could be assumed they progress in the whole volume of the material, not only in the surface layer. In order to overcome the problem, it is possible to use modern FTIR methods, which enable recording spectra of the surface layer only. One of the most prominent of these is attenuated total reflection FTIR (ATR-FTIR). Depending on the employed crystal, the technique scans spectra of the sample’s surface layer with a thickness from 0.5 to 5 μm, thus perfectly matching the typical thickness of one-layer biofim. In the instance of a biofilm presence, intensive signals corresponding to the proteins and polysaccharides involved should be expected. On the contrary, changes in the material are barely detectable because their signals are overlaid with those of biomolecules. The dominant signals result from the stretching vibration of C-O-C glykosidic bonds of polysaccharides in the 1200-900 cm-1 region, the signal at about 1650 cm-1 assigned to the stretching vibration of a C=O bond in amides, called amide I band, and a 1540 cm-1 signal from the deformation vibration of a N-H bond, named amide II band, both confirming the presence of proteins (Kansiz et al., 1999; Koutny et al., 2006). A signal at 2930 cm-1 originates from symmetric and asymmetric stretching vibrations of methyl groups, with another at 2875 cm-1 caused by symmetric and asymmetric vibrations of methylen groups. Methyl group signals are a sign of aliphatic amino acids in protein side chains, suspended ester groups in polysaccharides and terminal methyl groups in fatty acid molecules. The methylene signal may be also partially provoked by the presence of some constituents of proteins and polysaccharides. However, the contribution of methylene groups from fatty acids in lipids is dominant and the signal can be used to monitor their presence in a sample. When interpreting the ATR-FTIR spectra, it must be taken into account that signals from the sample material are usually also present and may eventually be intensive (Osiro et al., 2004). Also various signals can be slightly shifted with respect to literature data as a result of intramolecular or intermolecular interactions, eventually the maximum can appear shifted because of the superposition of more signals. The ATR-FTIR method and described signals in some cases can be used to monitor biofilm development on the surface of the material (Shmitt et al., 1998). The difficulty which is hard to overcome is that of obtaining quantitative or even semi-quantitative data. When inserting the sample into the instrument, it must be pressed sufficiently tightly against a measuring crystal so as to establish an optical contact. Since the surface of the sample is never perfectly smooth, there are no guarantees that the same or a well defined surface area of the sample will always come into optical contact. The alternative method for investigating a sample’s surface layer is photoacoustic infrared spectroscopy PAS-FTIR (Grodon et al., 1990). This technique makes it possible to obtain the spectra of compounds in a surface layer of depths from 1 to 100 μm, depending on the sample’s heat conductivity. The spectra are most likely to always contain signals from both biofilm and material.

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4.2. Nuclear Magnetic Resonance spectroscopy Nuclear Magnetic Resonance (NMR) spectroscopy is one of the major tools for chemist to analyze molecular structures; it is both a qualitative and quantitative method. More recently it was successfully applied to study microbial metabolisms including biodegradation of xenobiotics (Delort and Combourieu, 2001; Grivet et al., 2003). However, only a few papers report the use of this technique to assess biodegradation of non soluble polymers. A first approach consists in studying soluble low molecular compounds which result from the degradation of the polymer by classical liquid state NMR. This can be done by measuring directly the content of the incubation medium by in situ 1H NMR. The technique is very easy and straightforward, as it is just necessary to centrifuge the sample to eliminate solid particles and bacteria, add a reference and run NMR spectrum. This technique allowed monitoring the biodegradation of polyolephinic materials with prooxidant additives by Rhodococcus and Nocardia strains (Koutny et al. 2006). Even though the exact structures of all the metabolites were not identified, it could be used as a fingerprint (metabolic profile). An alternative is to extract short chain molecules from the solid material using an organic solvent, such as chloroform, dimethylsulfoxide or methanol, eventually make a rough separation by size exclusion chromatography, concentrate the extract and then run 1H or 13C NMR spectra. This approach was applied in the case of biodegradation studies of Mater-bi® and Eastar bio® materials (Massardier-Nageotte et al., 2006) and of lignin-based polymers (Xia et al., 2003). Some improvements could be made to determine the precise structure of degradation products, when present as complex mixtures, using more sophisticated 2D 1H and 13C NMR. This was successfully applied to study the biodegradation of wheat straw (e.g. Matulova et al., 2005). Finally solid-state NMR such as 13C-CPMAS (cross polarization magic angle spinning) could be in principle used to assess the structural changes of the solid material itself, this approach is particularly interesting because it gives also access to the crystallinity of the polymer (Matulova et al., 2005). However, two main limitations exist: first, NMR is not very sensitive, therefore larger changes have to occur to be detected; second, this technique gives information about the bulk material and is not adapted when molecular changes take place only at the surface.

4.3. Mass Spectroscopy Mass spectroscopy (MS) is also a very common analytical tool for chemists, and it was applied in some instances to analyze the structure of intermediates issued from polymer biodegradation. This technique is more sensitive than NMR but is not quantitative and usually needs to be coupled to a separation method such as gas chromatography for instance (GCMS). Readers can find examples in the literature for aliphatic-aromatic copolyesters (Witt et al., 2001), linear and branched poly(butylenes adipate) and poly(butylenes succinate) (Lindström et al., 2004) or polymer nanocomposites (Pandey et al., 2005). To improve the sensitivity, preconcentration of the sample can be made using solid phase extraction (Lindström et al., 2004).

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5. SIZE EXCLUSION CHROMATOGRAPHY The important method positively demonstrating changes in sample material is size exclusion chromatography, as used by many authors to also document biodegradation (Yamada-Onodera et al., 2001; Albertsson et al., 1995). Results can be displayed in the form of molecular weight distribution curves, or quantitatively formulated as weight or number average molecular weight, as well as a polydispersity index. It is possible to observe changes on distribution curves, signifying, for example, that low molecular weight substances were consumed preferentially (Kawai et al., 2004). Like other chromatographic techniques, the method shows usually slightly higher variability in its results between parallel runs, and it is more sensitive to minor differences in set-up method parameters. As a result of this, some minimal changes in distribution curves should not be interpreted as the sufficient evidence of biodegradation.

6. ESTIMATION OF BIOMASS CONTENT 6.1 Protein assays Protein determination can be classified as a classical approach to monitor biomass content inside a sample. Considering that proteins are an indispensable constituent of all living organisms, the resulting value can be used to asses the total weight of biomass present. The simple proportion is not encountered in instances of overproduction of some biomass constituent, for example extracellular polysaccharides, when protein content and total biomass content need not simply correspond. Among classical protein assays, the Bradford method (e.g. Orr, et al., 2004) and Lowry method (e.g. Orghan et al., 2000) have been used, to our knowledge, in biodegradation studies of water insoluble polymer materials. However, both assays have relatively poor sensitivity and enable detecting as much as 10 μg of proteins in ml. In this case, when attempting to observe the slow growth of biofilm on a tested material surface, such a level of sensitivity may prove insufficient. A marginally more sensitive method is the bicincholinic acid assay (Smith et al., 1985), with a limit of detection that could be, depending on the specific assay protocol, decreased to less than 1 μg. Nevertheless, every protein assay requires biofilm solubilization and the release of the protein into solution. This could be ensured with sample incubation in a solution of sodium hydroxide at an increased temperature. In the case of Lowry and bicincholinic acid methods, we can rely on the assumption that solubilization takes place during the assay incubation period as the employed reagents themselves are strongly alkaline. Long lists of interfering compounds have been described for each of the assays, and can also cause problems in this particular application. This is particularly true when the polymer was submitted to previous degradation, which makes the material capable of releasing a significant amount of low molecular compounds into its environment, some of which might interfere with a protein assay. Lowry and bicincholinic acid methods are not compatible with reducing substances; in the case of the Bradford assay, surface active substances must be avoided. In the experience of the authors, it follows that, in the case of oxidized polyethylene film, surface active substances were released from the material, in an amount even detectable

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by the naked eye, it was manifested by slight foaming after shaking. Interfering effects totally prevented use of the Bradford protein assay. The Lowry and bicincholinic acid methods were also not of assistance because of the very high background signal, which most probably originated from reducing substances leached from the oxidized polyethylene. Recently, new sensitive protein assays have been introduced based on the use of specific fluorescent dyes, an example of which is 3-(4-carboxybenzoyl)- quinoline-2-carboxaldehyde. This family of protein assays is of great interest thanks to their low detection limits of approximately 10 ng/ml and lower sensitivity to interfering agents, but there is no evidence of their use in the reviewed field at present.

6.2 Saccharide assays As an alternative to proteins, total biomass can be estimated from the content of polysaccharides. The classic method for this, frequently designated as the phenol-sulphuric assay (Dubois, 1956), is routinely employed for total extractable polysaccharide determination in complex matrixes like soils. Similarly, as was described for protein assays, this method is also made complicated by a number of interfering factors (Martens et al., 1993). The authors of the chapter are not aware of any studies in which polysaccharide determination is applied to investigate biodegradation of synthetic polymeric materials.

6.3 ATP and energetic status of microorganisms Adenosine triphosphate (ATP) plays a key role in the energetic metabolism of all living cells. Its cellular level is strictly controlled - a bacterial cell can contain, depending on its volume, about 2 × 1018 mol ATP (Lundin, 2000). Several means of bacterial ATP quantification have been described, but an especially useful and sensitive method is based on the use of the firefly enzyme system. This comprises enzyme luciferase and its substrate luciferin, where one of the products of ATP dependent reaction are photons, which subsequently can be very sensitively detected (Fig. 4).

Fig. 4. Reaction scheme of ATP determination with firefly enzymatic system. LH2, Luciferin; L, Oxyluciferin; E, Luciferase; ATP, Adenosine triphosphate; AMP, Adenosine monophosphate.

ATP determination must always be preceded by ATP extraction from a sample with a suitable extraction solution. As a consequence of the relatively aggressive effect of the

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extractant, the cells investigated must be destroyed and, at the same time, any ATP hydrolyzing enzymes present must be totally inactivated. The chief advantage of extraction is that ATP is extracted simultaneously from cells in suspension and from cells in biofilm on the surface of material. Thus, ATP determination enables the relatively simple and sensitive estimation of overall biomass in sample. This method was used in a biodegradation study of oxidized PE film with prooxidants with the defined bacterial strains of Rhodococcus rhodochrous and Nocardia astheroides (Koutny et al., 2006). An initial rapid growth of biomass was detected, probably as a result of extractable and easily degradable oxidation products followed by a decrease in the ATP level. After about 50 days of cultivation, the ATP level stabilized and maintained the same value to the end of the experiment, i.e. for another 150 days. Later, the described stable ATP level and related metabolic activity was interpreted as an indication of the ongoing but very slow biodegradation of the material. Even more information can be deduced from the ATP/ADP ratio (Rubia et al., 1986), or preferably from the so called ‘energetic charge’, the term for which is (ATP+0.5ADP)/(ATP + ADP + AMP), where ADP and AMP are amounts of adenosine diphosphate and adenosine monophosphate, respectively. Cited values are able to quantify the energetic state of microorganisms in a sample and show whether the biomass is metabolically active and growing or starving (Chapman et al., 1971). The parameters are also frequently used in the research of complex microbial communities, for example in soils (Dyckmans et al., 2003). Therefore, they could also prove useful to acquiring information on a sample’s biodegradation in soil or compost.

7. EVOLUTION OF CARBON DIOXIDE Monitoring CO2 production (anaerobic biodegradation with methanogenesis does not fall within the scope of the chapter) is often considered as the most demonstrative and straightforward quantitative method in biodegradation studies. Knowing the carbon content of a sample, which is relatively easy to obtain from elementary analysis, the quantity of CO2 obtained can be directly converted into the percentage of the sample mineralization. However, the real percentage of substrate biodegradation is always higher because a part of the sample is transformed into microbial biomass. The proportion of a sample’s carbon mineralization in the form of CO2 that becomes incorporated in the biomass can widely differ, depending as it does on experimental conditions and on the sample composition. In general, one broadly accepted idea is that the more oxidized material - a type of polysaccharides - is transformed in higher proportions to CO2, whereas for samples where carbon is present in lower oxidized states, the fraction of carbon build in biomass can even approach 50%. In its typical embodiment, a sample is mixed with a substrate containing active microorganisms and this mixture is placed into sealed column or flask, where it is flushed with air to supply the medium with oxygen and to expel carbon dioxide produced. Indeed, prior to this, incoming air must be humidified to prevent desiccation, and naturally occurring CO2 has to be removed by washing the air with an alkali hydroxide solution. The gas exhausted is also fed into an alkali hydroxide solution, where evolved CO2 is trapped. The carbon collected can be then determined in suitable intervals by volumetric titration. Mature compost can be used as the active microbial substrate and, in such cases, incubation is carried

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out at a temperature optimal for decomposing organic matter by thermophilic organisms, i.e. 58ºC, to simulate the composting process. Using soil instead of compost and incubation at normal temperatures allows for biodegradation of accidentally discarded material or biodegradation in a landfill to be investigated. Methodologies of both eventualities are described in detail in the form of the international standards ISO 14855 (compost) and ISO 17556 (soil). The described experiment’s set-up makes plain several serious inconveniences. For each sample, a relatively complicated installation with a number of tubes and joints must be constructed (Fig. 5).

Fig. 5. Diagram of apparatus employed in classical biometric study. Arrows indicate the direction of air flow. Evolving CO2 is absorbed in NaOH solution in gas washing bottles on the top of the figure.

Each sample must be tested in at least triplicate and, along with samples, at least two blank experiments must be run to establish endogenous production of CO2 evolved from the substrate alone. Another apparatus must be set aside for positive control to demonstrate sufficient microbial activity in the substrate. Considering the description above, one needs at least six flasks or columns to test one sample, naturally with all necessary tubing and gas washing bottles, resulting in a relatively complicated, difficult to maintain and sizeable system. The situation can become even more complicated if incubation is intended as the composting test at 58°C, making it necessity to control the temperature precisely. To conclude, this method is more suitable for samples with a rather high rate of biodegradation, and under circumstances of intense CO2 production. In such instances the experiment does not need to be carried out for an excessively long time, which is not always the case with synthetic polymer materials. An interesting modification of the technique was described by Chiellini et al. (2003). Samples in mixture with substrate were placed in sealed, wide-necked jars, inside of which

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beakers with CO2 trapping solution were placed; internal atmospheric circulation was only ensured by diffusion. Trapped carbon was determined by a volumetric titration. This configuration was successfully used in a study on oxidized polyethylene biodegradation that spanned more than one year. The passive gas exchange could be considered among the inconveniences of such a set-up as there was no control mechanism to monitor if the supply of oxygen for sample decomposition was sufficient; and there is a risk that some CO2 may elude detection).

Fig. 6. Cultivation flask for gas chromatography determination of CO2 production. a, sampling septum; b, inlet for aeration; c, needle conducting the incoming air to the bottom of the flask; b outlet for aeration.

Of particularly advantage is a modification where CO2 produced is analyzed by gas chromatography in a head space atmosphere over a sample substrate mixture (Drimal et al.,

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2006). In parallel, oxygen concentration can also be monitored to provide a control so that samples do not suffer from hypoxia. Incubation is done in sealed flasks equipped with septum on stoppers, which are eventually fitted with another two valves for aeration between measuring cycles (Fig. 6). Mixing of the content can be done at any time by shaking the flask. Head space gas was sampled through the septum with a gas-tight syringe and was then injected manually into a GC instrument. Sampling can be operatively adapted to the actual CO2 production. If the residual concentration of oxygen approaches 30% (v/v) of oxygen concentration in the air, the internal atmosphere must be exchanged for fresh air. An important issue complicating the situation, especially during long term incubations of slow degradable materials, can be endogenous CO2 production by the substrate alone, which is particularly significant when compost is used as the substrate and incubation is run under conditions for biodegradation by thermophilic microorganisms. Certain assistance could be found by replacing a significant part of the substrate, for example one half, with inert material (Drimal et al., 2007), a commonly used alternative is perlite. Either soil or compost then almost entirely adopt the role of microbial inoculum. To further reduce background CO2, other authors proposed a total exchange of compost with clay mineral vermiculite following its activation (Bellia et al., 1999). The activation process is based on incubating vermiculite with a small amount of compost added as a microbial inoculum, and the provision of an easily utilizable carbon source (cellulose, starch, nutrient broth). Taken as a whole, the idea can work well if biodegradation of similar material, like the one used for activation, is investigated. Should this not be the case, problems can be encountered. Activated vermiculite will contain only fast growing microorganisms multiplied on easily utilizable carbon sources during the activation; the microorganisms will not support biodegradation of others slowly and/or more resistant materials requiring specialized microorganisms, which can in fact be suppressed during the activation process. Results of biodegradation experiments also strongly depend on the quality of the employed inoculum (Mezzanotte et al., 2005). All the described monitoring techniques for CO2 production can be adapted for experiments with defined microbial strains in aqueous media, or in moistened solid inert media, such as perlite.

8. STANDARDIZATION OF METHODS Recently, biodegradable materials have become increasingly favored by legislative acts in a growing number of countries. A given material can be classified as biodegradable via a process requiring standardized procedures and clearly defined criteria, which is the role of international standards. However, biodegradable materials constitute a broad group of materials for various purposes. These can differ significantly in conditions where sufficient biodegradation might occur, as well as in the time scale of biodegradation. Over the last few years, existing standards were often subjected to criticism, hence effort has been made to revise and consolidate them (Krzan et al., 2006). Most of the standards applicable to water insoluble plastic materials had evolved from procedures initially developed for evaluating compounds that biodegrade relatively easily in an aqueous

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environment. Consequently, criteria are relatively strict and given limits must be achieved in an unreasonably short time. As regards other standards, the criteria are not fully relevant or significant to biodegradation, such as sample disintegration or observation of microorganism growth on a sample surface. On the other hand, a credible standard must also take into account the fact that some producers may deliberately try to evade criteria, for example by using an easily biodegradable filler in an unbiodegradable matrix and then attempt to introduce the resulting blend on the market as biodegradable.

9. CONCLUSION Constant search for new methods and their introduction in to the polymer material biodegradability testing evidences need for continuing research in the field. Till today results from none of the methods described can be taken as an absolute. The final statement on a material’s biodegradability should always be made after consulting results from more experimental approaches. Thoroughly considered, it is virtually impossible to classify a material definitively as non-biodegradable. This is because a possibility always exists that the experiment was done under unfavorable conditions for the given type of material, or that the specific necessary microorganisms were not present, or simply that the biodegradation is too slow to be detected by methods employed. In contrast, a lot of studies can be found in literature in which authors claim biodegradability on the basis of very vague or insignificant observations.

REFERENCES Albertsson, A.-C., Barenstedt, C., Karlsson, S. and Lindberg, T. (1995). Degradation product pattern and morphology changes as means to differentiate abiotically and biotically aged degradable polyethylene, Polymer, 36, 3075–3083. Arnaud, R., Dabin, P., Lemaire, J., Al-Malaika, S., Chohan, S., Coker, M., Scott, G., Fauve, A. and Maaroufi, A. (1994). Photooxidation and biodegradation of commercial photodegradable polyethylenes. Polymer Degradation and Stability, 46(2), 211-224. Bellia, G., Tosin, M., Floridi, G. and Degli-Innocenti, F. (1999) Activated vermiculite, a solid bed for testing biodegradability under composting conditions. Polymer Degradation and Stability, 66(1), 65-79. Bonhomme, S., Cuer, A., Delort, A-M., Lemaire, J., Sancelme, M. and Scott, G. (2003). Environmental biodegradation of polyethylene. Polymer Degradation and Stability, 81(3). 441-452. Chapman, A., G., Fall, L., and Atkinson, D., E. (1971). Adenylate Energy Charge in Escherichia coli During Growth and Starvation. Journal of bacteriology 108(3), 10721086.

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In: Environmental Biodegradation Research Focus Editor: B. Y. Wang, pp. 259-271

ISBN: 978-1-60021-904-7 © 2007 Nova Science Publishers, Inc.

Chapter 10

BIODEGRADABLE SYNTHETIC OCTACALCIUM PHOSPHATE BONE SUBSTITUTE Takahisa Anada1, Hideki Imaizumi2, Shinji Kamakura3 and Osamu Suzuki1∗ 1

Division of Craniofacial Function Engineering (CFE), Tohoku University Graduate School of Dentistry, Japan 2 Department of Orthopedic Surgery, Osaki Citizen Hospital, Furukawa, Osaki, Japan. 3 Division of Clinical Cell Therapy, Department of Translational Research, Center for Translational and Advanced Animal Research (CTAAR), Tohoku University School of Medicine, Aoba-ku, Sendai, Japan

ABSTRACT Octacalcium phosphate (OCP) has been advocated to be a precursor of biological apatite crystals in bones and teeth. In fact, several studies using physical techniques demonstrated that OCP is involved as a transitory intermediate phase to biological apatite crystals in enamel, dentine and bone. Our previous studies demonstrated that synthetic OCP facilitates bone regeneration, compared to synthetic hydroxyapatite (HA), including non-sintered stoichiometric or non-stoichiometric HA, and sintered HA ceramic, when implanted in murine and rabbit bone defects. Synthetic OCP can be replaced with newly formed bone in conjunction with its simultaneous biodegradation. Crystallographic OCPapatite conversion of the implanted OCP advances gradually during the bone regeneration. OCP is known to be a thermodynamically less stable salt under physiological condition than HA and β-tricalcium phosphate (β-TCP); the latter is a well known biodegradable bone substitute ceramic. The biodegradability predicted from the solubility isotherm is actually reproduced in the implantation of these synthetic calcium phosphate compounds into bone defects. OCP-apatite conversion induces various ∗

Corresponding author: Professor Osamu Suzuki, Ph.D. 1Division of Craniofacial Function Engineering (CFE) Tohoku University, Graduate School of Dentistry, 4-1 Seiryo-machi, Aoba-ku, Sendai 980-8575, Japan; TEL: +81-22-717-7635, FAX: +81-22-717-7637; E-Mail: [email protected]

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Key words: Bone substitute material, octacalcium phosphate (OCP), bone formation, biodegradation

1. BIODEGRADABLE BIOMATERIALS FOR BONE TISSUE ENGINEERING Transplantation of autologous bone grafts is the gold standard for orthopedic, oral and plastic surgical procedures. However, it has disadvantages especially limited availability and the harvest procedure involves high donor site morbidity [1]. Alternative to autologous transplant materials are currently being investigated. A variety of materials have been used for regeneration of bone tissues. Biodegradable polymer matrices are widely studied for bone tissue engineering applications [2]. There are two types of biodegradable polymers. The naturally derived materials are one category, such as proteins (collagen, fibrin) or polysaccharides (starch, alginate, chitosan, hyaluronic acid). Synthetic polymers are the second category, including poly(lactic acid) (PLA), poly(glycolic acid ) (PGA), and their copolymers (PLGA). However, polymeric materials alone can not meet all the requirements for bone regeneration. Some of those materials implanted are encapsulated by a fibrous tissue and do not adhere to bone directly because of lack of bioactivity [3]. To resolve these problems, bioactive materials, such as calcium phosphate, have been proposed as bone substitute.

2. CALCIUM PHOSPHATE BONE SUBSTITUTE Table 1 shows calcium phosphate compounds associated with biominerals in vertebrates. The Ca/P molar ratio is one of the indices used to identify particular calcium phosphate compound. The prototype for the mineral in bones and teeth is usually considered to be basic calcium phosphate HA [4, 5]. The biomineral displays better crystallinity and higher molar ratio with mineral development [4, 5]. The final biological apatite crystals are constituted by poorly crystalline HA with low Ca/P molar ratio, i.e. Ca-deficient HA, containing foreign ions such as carbonate and fluoride. Synthetic inorganic calcium phosphates, such as hydroxyapatite (HA), and betatricalcium phosphate (β-TCP), have been widely used as a bone substitute in orthopedic and dental surgery [6-8]. HA and β-TCP are recognized as biocompatible and osteoconductive. They have somewhat different ratios of calcium and phosphorus, however, the material properties are essentially mimic closely the inorganic phase of bone, which constitutes 60-70 % of human bone [9]. These materials can act as a scaffold for osteogenic cells during bone regeneration process. Sintered HA is considered to be remained undissolved [10, 11]. On the contrary, several lines of evidence demonstrated that β-TCP slowly degrades in vivo, and eventually gets completely replaced with host bone [7, 12]. It is considered that the solubility of calcium phosphate ceramics affects osteoclast resorption activity and the resorption pattern

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[13]. These in vitro observations confirm in vivo data on comparative degradation of calcium phosphate materials. Selection of appropriate materials with the biodegradation is required to apply for particular clinical situation. Table 1 Calcium phosphate compounds relating to hydroxyapatite Calcium phosphate

Abbreviation

Chemical formula

Ca/P molar ratio

dicalcium phosphate dihydrate

DCPD

CaHPO4･2H2O

1.00

octacalcium phosphate

OCP

Ca8H2(PO4)6･5H2O

1.33

tricalcium phosphate

TCP

Ca3(PO4)2

1.50

amorphous calcium phosphate

ACP

Ca3(PO4)2･nH2O

1.50

hydroxyapatite

HA

Ca10(PO4)6(OH)2

1.67

3. OCTACALCIUM PHOSPHATE Although there is still a controversy about the chemical nature of the first biomineral formed [14, 15], OCP has been suggested to be a precursor to HA because in vitro crystal growth studies showed that OCP can be present as an intermediate during formation of HA [16-18]. Multiple lines of evidence obtained in vitro and in vivo support the involvement of OCP in the initial formation of dentin, enamel and bone minerals [16, 17, 19-23]. It is most likely that OCP precursors, if they exist, may have a transitory nature during the growth of biological apatite and can only be detected in larger crystals because of their slow kinetics of transformation [19]. Conversion of synthetic OCP into HA has been investigated in various physiological media, such as simulated body fluid [24] ultrafiltered human serum [25], and by the murine tissue implantation [26-29]. It has been shown that the converted apatite from OCP in vitro physiological conditions was a Ca-deficient HA which has a chemical composition with lower Ca/P molar ratio and higher acid phosphate content [30], as observed in biological crystals [31]. The Ca-deficient HA formed via OCP contained small amount of OCP as residual inclusion within the formed apatite [32, 33] and further retained the original platy morphology of OCP [30, 32]. The osteoconductive characteristic of synthetic OCP has been found in mouse calvarial bone responses 10 years ago [28, 29]. However, the study to elucidate biological significance of OCP, if implanted in bone, is still under way.

4. OCP AS BONE SUBSTITUTE It has been reported that implanted OCP could be bone substitute in the body [28, 29, 3437]. Figure 1 shows radiographs and histological examination of rat skull defects treated with OCP implantation at 4 (a, d), 12 (b, e), and 24 (c, f) weeks [34]. At week 4, isolated granulous radiopacities are scattered throughout the defect (a). New bone is formed in two fashions; from the defect margin of the calvaria and on the implanted OCP away from the margin.

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Osteoblasts are seen on newly formed bone around the implanted OCP (d). At week 12, larger radiopacities are formed by amalgamation of the isolated radiopacities (b). Implanted OCP is surrounded by multinucleated giant cells (MNGCs) (e). At week 24, the amalgamated radiopacities merge with the radiopacities from the margin of the bone defect and become plate-like. As seen in Figure 1-f, the defect is almost filled with newly formed bone and the remaining OCP. Figure 2 shows histological examination with hematoxylin-eosin staining by serial sections on a rat skull defect treated with the implantation of OCP in week 4 [35]. Observation of the serial sections showed that some of the new bones on the implanted OCP were formed away from the defect margin of the parietal bone and osteocytes were observed in the newly formed bone. These results indicate that the implanted OCP serves as a core for initiating bone regeneration. The newly formed bone matrix around the OCP implant had phenotypes of type I collagen [38] and osteocalcin [35].

Figure 1: Radiographs and histological examination treated with OCP implantation at 4 (a, d), 12 (b, e), and 24 (c, f) weeks. (a) At week 4, isolated granulous radiopacities are scattered throughout the defect. (b) At week 12, larger radiopacities are formed by amalgamation of the isolated radiopacities. (c) At week 24, the amalgamated radiopacities merge with the radiopacities from the margin of the bone defect and become plate-like. (d) At week 4, osteoblasts (open arrows) are seen on newly formed bone (B) around the implanted OCP (*). It is surrounded by multinucleated giant cells (open arrowheads). (e) At week 12, implanted OCP (*) is surrounded by newly formed bone (B). (f) At week 24, the defect is almost filled with newly formed bone (B) and the remaining OCP (*). Bars = 3 mm (a ~ c), 100 µm. (d ~f). (Reprinted with permission and reorganized from Kamakura S. et al.: Implantation of octacalcium phosphate (OCP) in rat skull defects enhances bone repair. J Dent Res78: 1682-1687, 1999, The International and American Associations for Dental Research)

Bone regeneration of implanted OCP, β-TCP, and HA were analyzed histomorphometrically, after long-term implantation (6 months) into an experimental cranial defect in rats [37]. Figure 4A shows statistical analysis showed that the percentage of new bone in the defect (n-Bone%) ± SD within the OCP, β-TCP, and HA groups is 63.3 ± 7.52, 42.8 ± 13.6, and 24.2 ± 7.82, respectively. The difference of mean value of the n-Bone% between OCP-treated and other two groups, and that of the n-Bone% between β-TCP-treated group and HA-treated group are significant. These results suggest that implanted OCP enhances bone formation compared with HA or β-TCP. Other investigators also reported that

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OCP coating on metallic implants showed osteoconductive properties [39-42] and in some cases osteoinductive [43].

Figure 2: Histological examination by serial sections (1-9) on a skull defect treated with the implantation of OCP in week 4. New bone (B) on the implanted OCP (*) shows no continuity with the components that define the defect. Osteocytes are observed in the newly formed bone (4-7). Implanted OCP is no longer seen in (1), (2) and (3). *: implants, B: newly formed bone, Bars = 100 µm. (Reprinted with permission and reorganized from Kamakura S. et al.: Implantation of octacalcium phosphate nucleates isolated bone formation in rat skull defects. Oral Dis 7: 259-265, 2001, Blackwell Publishing)

5. BIODEGRADABILITY OF OCP The biodegradation of calcium phosphate biomaterials is generally assumed to take place by solution-driven and cell-mediated processes [44] and to be influenced by the experimental conditions, such as experimental models, implantation sites, and animal species [45]. It is known that OCP is physicochemically resorbable more than HA or β-TCP under the physiological condition [16]. We investigated whether OCP implantation enhances bone remodeling concomitant with OCP degradation when implanted cranial defect of rats [37] or intramedullary in a rabbit femur [10]. Figure 3 shows comparison of histomorphometrically resorption of the implanted OCP, HA, and β-TCP, which were kept in the experimental cranial defect of rats in 6 months. In the OCP-treated rats, the remaining implants in the defect were surrounded by newly formed bone without intervening cellular components (Figure 3 a, d) and no remaining OCP was seen in the connective tissue. In the β-TCP-treated rats, the remaining implants in the defect were seen both in the newly formed bone and in the connective tissue. The remaining β-TCP in bone was directly in contact with the newly formed bone (Figure 3e). In the HAtreated rats, the remaining implants in the defect were both in the newly formed bon and in the connective tissue. The remaining HA in bone was directly in contact with the newly formed bone (Figure 3f). Figure 4B shows histomorphometrical examination of remaining implants in the defect (r-Imp%) caused by OCP, β-TCP, and HA. The percentage of

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remaining implants in the defect (r-Imp%) ± SD within OCP, β-TCP, and HA group is 6.58 ± 1.46, 23.4 ± 6.82, and 39.0 ± 6.86, respectively. The r-Imp% of the OCP-treated group was significantly lower than that of the β-TCP- or HA-treated group and the r-Imp% of the βTCP-treated group was significantly lower than that of the HA-treated group. It suggests that the implanted OCP is more resorbable than the β-TCP or the HA, and the implanted β-TCP is more resorbable than HA. It has also been reported that the resorption pits were observed on preincubated OCP disks but not on HA and β-TCP disks, when incubated in osteoclastic cell culture [46].

Figure 3: Histological examination treated with OCP (a, d), β-TCP (b, e), and HA (c, f) at 6 months (a, d) The created defect is mainly filled with newly formed bone (B) which is compact and surrounds directly the implanted OCP (*) without intervening cellular components. (b, e) The newly formed bone (B) is compact and directly contacts the implanted β-TCP (*) without intervening cellular components. The remaining βTCP (*) in the defect surrounds by the newly formed bone (B) and the connective tissue. (c, f) The newly formed bone (B) directly contacts the implanted HA (*) without intervening cellular components. The remaining HA (*) in the defect surrounds by the newly formed bone (B) and the connective tissue. ▼: defect margin, *: implants, B: newly formed bone, Bars = 1 mm (a, c, e), 200 µm (b, d, f). (Reprinted with permission and reorganized from Kamakura S. et al.: Implanted octacalcium phosphate is more resorbable than β-tricalcium phosphate and hydroxyapatite. J Biomed Mater Res 59: 29-34, 2002, John Wiley & Sons, Inc.)

Figure 5 shows histological section of OCP, when OCP was implanted intramedullary in a rabbit femur for 3 weeks [10]. MNGCs were attached to the bare OCP surface. When implanted into rat bone marrow, the implanted OCP and HA were surrounded by MNGCs [36]. MNGCs on the implanted OCP developed a ruffled border-like structure and clear zonelike structure, showing ultrastructural characteristics with osteoclasts. Implanted OCP was vigorously resorbed in contact with the ruffled border-like structure. MNGCs on the implanted HA develop a clear zone-like structure, whereas no ruffled border was seen. The surface of the implanted HA associated with MNGC was smooth and not resorbed.

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A

265

B

Figure 4: Histomorphometrical examination of newly formed bone in the defect (n-Bone%) (A) and remaining implants in the defect (r-Imp%) (B) caused by OCP, β-TCP, and HA. (A) The n-Bone% ± SD within the OCP, β-TCP, and HA groups is 63.3 ± 7.52, 42.8 ± 13.6, and 24.2 ± 7.82, respectively. The difference of mean value of the n-Bone% between OCP-treated and other two groups, and that of the nBone% between β-TCP-treated group and HA-treated group are significant. (B) The r-Imp% ± SD within OCP, b-TCP, and HA group is 6.58 ± 1.46, 23.4 ± 6.82, and 39.0 ± 6.86, respectively. The difference of mean value of the r-Imp% between OCP-treated and other two groups and that of the r-Imp% between βTCP-treated group and HA-treated group are significant. Data are means ± SD of five specimens. Statistical analysis is performed using an unpaired t test. *: p < 0.05 (Reprinted with permission and reorganized from Kamakura S. et al.: Implanted octacalcium phosphate is more resorbable than β-tricalcium phosphate and hydroxyapatite. J Biomed Mater Res 59: 29-34, 2002, John Wiley & Sons, Inc.)

Figure 5: Undecalcified histological sections treated with OCP at 3 weeks. Mature appositional bone (trabecular bone) was formed around OCP, and MNGCs were directly attached to the OCP surface. *: implants, Bar = 50 μm. (Reprinted with permission and reorganized from Imaizumi, H. et al.: Comparative study on osteoconductivity by synthetic octacalcium phosphate and sintered hydroxyapatite in rabbit bone marrow. Calcif Tissue Int 78: 45-54, 2006, Springer Science + Business Media, Inc.)

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Figure 6 shows the results obtained by histomorphometric analysis, when OCP was implanted intramedullary in a rabbit femur, compared to sinter HA ceramic [10]. Appositional bone formation (MS/IMPS) was significantly higher on OCP at 2 (P < 0.05) and 3 (P < 0.01) weeks than on HA. MS/IMPS was significantly lower on OCP than on HA at 12 weeks (P < 0.05). MS/IMPS was 45.4 ± 3.7 % for OCP and 21.6 ± 6.5 % for HA at 2 weeks. MS/IMPS was 44.5 ± 2.8 % for OCP and 19.2 ± 4.5 % for HA at 3 weeks (Figure 6B). There was no significant difference in MNGCS/IMPS until 4 weeks; it became significantly higher on OCP (27.8 ± 3.9 %) than on HA (14.9 ± 1.5 %) at 8 weeks (P < 0.05) (Figure 6C). Figure 7 shows radiographs of rabbit femurs with OCP and HA at 4 and 12 weeks. Radiopacity derived from the OCP implant was evident at 4 weeks but became obscure at 12 weeks (Figure 7a, b), whereas the HA implant remained almost unchanged (Figure 7c, d). These results indicate that OCP granules were biodegraded, whereas those of HA were not.

A

B

C

Figure 6: Rationale for determining appositional bone formation defined by MS/IMPS and MNGCS/IMPS (A). Estimation of MS/IMPS (%, B), and MNGCS/IPMS (%, C) on implants using histological sections. Each value was expressed as the mean ± SD of n = 5. (A) Appositional bone formation (MS/IMP) was defined as the ratio of the length of mineralized bone formed and attached on the implant surface (MS: A to B plus C to D) to the length of the implant circumference (IMPS: A to A). MNGCS/IMPS was expressed as the ratio of the length of the implant surface that MNGCS attacked (MNGCS: A to F plus E to D) to the length of the implant circumference (IMPS: A to A) (B) MS/IMPS of OCP was significantly higher than that of HA at 2 and 3 weeks and lower than that of HA at 12 weeks (*P < 0.05, **P < 0.01). (C) MNGCS increased significantly on the OCP implant at 8 weeks (*P < 0.05). (Reprinted with permission and reorganized from Imaizumi, H. et al.: Comparative study on osteoconductivity by synthetic octacalcium phosphate and sintered hydroxyapatite in rabbit bone marrow. Calcif Tissue Int 78: 45-54, 2006, Springer Science + Business Media, Inc.)

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Figure 7: Radiographs of a rabbit femur including the site where the OCP or HA ceramic was implanted. (a, b) OCP implantation. (c, d) HA ceramic implantation. (a, c) Four weeks, (b, d) 12 weeks. Note that radiopacity, derived from the OCP implant and newly formed mineralized bone, was evident at 4 weeks (a), whereas it became obscure at 12 weeks (b), indicating biodegradation of the OCP implant and resorption of newly formed bone. Radiopacity remained almost unchanged for HA implantation (c, d). Bars = 9 mm. (Reprinted with permission and reorganized from Imaizumi, H. et al.: Comparative study on osteoconductivity by synthetic octacalcium phosphate and sintered hydroxyapatite in rabbit bone marrow. Calcif Tissue Int 78: 45-54, 2006, Springer Science + Business Media, Inc.)

6. OSTEOGENIC PROPERTY OF OCP OCP are potential scaffold materials in the tissue-engineering field as described above. However, the molecular mechanism of the induction of osteoconductive characteristics by OCP has not been fully elucidated. In a previous report, we investigated the influence of OCP crystal on proliferation and differentiation of osteoblastic cells in vitro [26, 27]. Cell culture dishes were coated with the slurry of synthetic OCP, OCP hydrolyzate and stoichiometric HA. A cell line established from mouse bone marrow stromal ST2 cells was inoculated into the dishes coated with the calcium phosphate compounds. Cell growth was determined by

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counting the cell numbers under a fluorescent microscope after nuclear staining using a fluorescent dye, DAPI. Although ST2 cells proliferated well on stoichiometric HA as well as on the control dish, they proliferated slower on OCP than that on stoichiometric HA or OCP hydrolyzate. On the other hand, their differentiation to osteoblasts was promoted on OCP at 20 days culture. The lower proliferation capacity and promotion of differentiation of ST2 cells on OCP in vitro may be related to the surface property of the crystal. Other groups also reported that OCP enhanced the differentiation from rat bone marrow stromal cells to osteoblastic cells [47, 48]. In these studies, mRNA of osteocalcin was upregulated for 24 days culture, which may correspond to the stimulation the mature stage of cell differentiation. The hydrolysis reaction of coated OCP into apatite was also achieved during cell culture incubation as well as in vivo conversion [28, 29]. The microenvironment adjacent to the OCP crystal should be somewhat acidic as expected from the hydrolysis reaction of OCP due to the release of phosphate ions [16, 26] and would affect the growth and/or function of cells in vitro. In order to establish a link between OCP activity in vitro and in vivo, the mechanism underlying the surface reaction through biomolecule adsorption and OCP hydrolysis during cell culture is currently being studied.

7. CONCLUSION Interest in scaffolds for bone tissue repair is continuously growing. Synthetic calcium phosphates have become more popular for bone repair in orthopedic and dental surgery. There is a general consensus that scaffolds for bone regeneration should be biocompatible. Furthermore, it is preferably to be bioresorbable or biodegradable, because bone formation by osteoblasts is associated with bone resorption by osteoclasts during the bone-remodeling process. The biodegradation of calcium phosphate biomaterials is supposed to take place by solution-drive and cell-mediated processes. OCP has been shown to be converted into an apatite structure when implanted in murine calvarial bone [27-29], to enhance bone regeneration more than synthetic HA [10, 37], and to degrade faster than biodegradable βTCP [37]. OCP probably offers advantage of bone remodeling process and providing formation of strong bone-material interface. Further investigations are needed to fully understand the bone regeneration and biodegradation mechanism by the implantation of synthetic OCP. However, it is expected that synthetic OCP can be used as an implant material in the field of orthopedic and craniofacial bone regeneration due to its properties of enhancing bone formation.

REFERENCES [1] Goulet JA, Senunas LE, DeSilva GL, Greenfield ML. Autogenous iliac crest bone graft. Complications and functional assessment. Clin Orthop 1997;339:76-81. [2] Burg KJL, Porter S, Kellam JF. Biomaterial developments for bone tisse engineering. Biomaterials 2000;21:2347-59.

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[3] Kokubo T, Kim HM, Kawashita M. Novel bioactive materials with different mechanical properties. Biomaterials 2003;24:2161-75. [4] Aoba T. Recent observations on enamel crystal formation during mammalian amelogenesis. Anat Rec 1996;245:208-18. [5] Kim HM, Rey C, Glimcher MJ. Isolation of calcium-phosphate crystals o bone by nonaqueous methods at low temperature. J Biomed Mater Res 1995;10:1589-601. [6] Bucholz RW. Nonallograft osteoconductive bone graft substitutes. Clin Orthop 2002;395:44-52. [7] LeGeros RZ. Properties of osteoconductive biomaterials: Calcium phosphates. Clin Orthop 2002;395:81-98. [8] Ogose A, Hotta T, Kawashima H, Kondo N, Gu W, Kamura T, Endo N. Comparison of hydroxyapatite and beta-tricalcium phosphate as bone substitutes after excision of bone tumors. J Biomed Mater Res B Appl Biomater 2005;72:94-101. [9] Steffen T, Stoll T, Arvinte T, Schenk RK. Porous tricalcium phosphate and transforming growth factor used for anterior spine surgery. Eur Spine J 2001;10(Suppl 2):S132-40. [10] Imaizumi H, Sakurai M, Kashimoto O, Kikawa T, Suzuki O. Comparative study on osteoconductivity by synthetic octacalcium phosphate and sinterd hydroxyapatite in rabbit bone marrow. Calcif Tissue Int 2006;78:45-54. [11] Uchida A, Araki N, Shinto Y, Yoshikawa H, Kurisaki E, Ono K. The use of calcium hydroxyapatite ceramic in bone tumour surgery. J Bone Joint Surg 1990;72-B:298-302. [12] Gaasbeek RDA, Toonen HG, van Heerwaarden RJ, Buma P. mechanism of bone incorporation of beta-TCP bone substitute in open wedge tibial osteotomy in patients. Biomaterials 2005;26:6713-9. [13] Yamada S, Heymann D, Bouler JM, Daculsi G. Osteoclastic resorption of calcium phosphate ceramics with different hydroxyapatite/beta-tricalcium phosphate ratios. Biomaterials 1997;18:1037-41. [14] Boskey AL. Biomineralization: Conflicts, challenges, and opportunities. J Cell Biochem Suppl 1998;30-31:83-91. [15] Grynpas MD, Omelon S. Transient precursor strategy or very small biological apatite crystals? Bone 2007;41:162-4. [16] Brown WE, Mathew M, Tung MS. Crystal chemistry of octacalcium phosphate. Prog Cryst Growth Charact 1981;4:59-87. [17] Mann S. Biomineralization. New York: Oxford University Press 2001:38-67. [18] Meyer JL, Eanes ED. A thermodynamic analysis of the secondary transition in the spontaneous precipitation of calcium phosphate. Calcif Tissue Res 1978;25:209-16. [19] Bodier-Houlle P, Steuer P, Voegel JC, Cuisinier FJ. First experimental evidence for human dentine crystal formation involving conversion of octacalcium phosphate to hydroxyapatite. Acta Crystallogr D Biol Crystallogr 1998;54(Pt 6 Pt 2):1377-81. [20] Diekwisch TG, Berman BJ, Gentner S, Slavkin HC. Initial enamel crystals are not spatially associated with mineralizaed dentine. Cell Tissue Res 1995;279:149-67. [21] Iijima M, Tohda H, Y M. Growth and structure of lamellar mixed crystals of octacalcium phosphate and apatite in a model system of enamel formation. J Cryst Growth 1992;116:319-26. [22] Miake Y, Shimoda S, Fukae M, Aoba T. Epitaxial overgrowth of apatite crystals on the thin-ribbon precursor at early stages of porcine enamel mineralization. Calcif Tissue Int 1993;53:249-56.

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[23] Siew C, Gruninger SE, Chow LC, Brown WE. Procedure for the study of acidic calcium phosphate precursor phases in enamel mineral formation. Calcif Tissue Int 1992;50:1448. [24] Lu X, Leng Y. Theoretical analysis of calcium phosphate precipitation in simulated body fluid. Biomaterials 2005;26:1097-108. [25] Eidelman N, Chow LC, Brown WE. Calcium phosphate phase transformations in serum. Calcif Tissue Int 1987;41:18-26. [26] Suzuki O, Kamakura S, Katagiri T. Surface chemistry and biological responses to synthetic octacalcium phosphate. J Biomed Mater Res Appl Biomater 2006;77B:201-12. [27] Suzuki O, Kamakura S, Katagiri T, Nakamura M, Zhao B, Honda Y, Kamijo R. Bone formation enhanced implanted octacalcium phosphate involving conversion into Cadeficient hydroxyapatite. Biomaterials 2006;27:2671-81. [28] Suzuki O, Nakamura M, Miyasaka Y, Kagayama M, Sakurai M. Bone formation on synthetic precursors of hydroxyapatite. Tohoku J Exp Med 1991;164:37-50. [29] Suzuki O, Nakamura M, Miyasaka Y, Kagayama M, Sakurai M. Maclura pomifera agglutinin-binding glycoconjugates on converted apatite from synthetic octacalcium phosphate implanted into subperiosteal region of mouse calvaria. Bone Miner 1993;20:151-66. [30] Suzuki O, Yagishita H, Yamazaki M, Aoba T. Adsorption of bovine serum albumin onto octacalcium phosphate and its hydrolyzates. Cells Mater 1995;5:45-54. [31] Shimoda S, Aoba T, Moreno EC. Changes in acid-phosphate content in enamel mineral during porcine amelogenesis. J Dent Res 1991;70:1516-23. [32] Iijima M, Nelson DG, Pan Y, Kreinbrink AT, Adachi M, Goto T, Moriwaki Y. Fluoride analysis of apatite crystals with a central planar OCP inclusion: concerning the role of Fions on apatite/OCP/apatite structure formation. Calcif Tissue Int 1996;59:377-84. [33] Nelson DG, Barry JC. High resolution elctron microscopy of nonstoichiometric apatite crystals. Anat Rec 1989;224:265-76. [34] Kamakura S, Sasano Y, Homma H, Suzuki O, Kagayama M, Motegi K. Implantation of octacalcium phosphate (OCP) in rat skull defects enhances bone repair. J Dent Res 1999;78:1682-7. [35] Kamakura S, Sasano Y, Homma H, Suzuki O, Kagayama M, Motegi K. Implantation of octacalcium phosphate nucleates isolated bone formation in rat skull defects. Oral Dis 2001;7:259-65. [36] Kamakura S, Sasano Y, Homma-Ohki H, Nakamura M, Suzuki O, Kagayama M, Motegi K. Multinucleated giant cells recruited by implantation of octacalcium phosphate (OCP) in rat bone marrow share ultrastructural characteristics with osteoclasts. J Electron Microsc 1997;46:397-403. [37] Kamakura S, Sasano Y, Shimizu T, Hatori K, Suzuki O, Kagayama M, Motegi K. Implanted octacalcium phosphate is more resorbable than beta-tricalcium phosphate and hydroxyapatite. J Biomed Mater Res 2002;59:29-34. [38] Sasano Y, Kamakura S, Nakamura M, Suzuki O, Mizoguchi I, Akita H, Kagayama M. Subperiosteal implantation of octacalcium phosphate (OCP) stimulates both chondrogenesis and osteogenesis in the tibia, but only osteogenesis in the parietal bone of a rat. Anat Rec 1995;242:40-6.

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[39] Barrere F, van der Valk CM, Dalmeijer RA, Meijer G, Van Blitterswijk CA, de Groot K, Layrolle P. Osteogenecity of octacalcium phosphate coatings applied on porous metal implants. J Biomed Mater Res A 2003;66:779-88. [40] Barrere F, van der Valk CM, Dalmeijer RA, Van Blitterswijk CA, de Groot K, Layrolle P. In vitro and in vivo degradation of biomimetic octacalcium phosphate and carbonate apatite coatings on titanium implants. J Biomed Mater Res A 2003;64:378-87. [41] Bigi A, Bracci B, Cuisinier F, Elkaim R, Fini M, Mayer I, Mihailescu IN, Socol G, Sturba L, Torricelli P. Human osteoblast response to pulsed laser deposited calcium phosphate coatings. Biomaterials 2005;26:2381-9. [42] Dekker RJ, de Bruijn JD, Stigter M, Barrere F, Layrolle P, CA. vB. Bone tissue engineering on amorphous carbonated apatite and crystalline octacalcium phosphatecoated titanium discs. Biomaterials 2005;26:5231-9. [43] Habibovic P, Li J, van der Valk CM, Meijer G, Layrolle P, van Blitterswijk CA, de Groot K. Biological performance of uncoated and octacalcium phosphate-coated Ti6Al4V. Biomaterials 2005;26:23-36. [44] Lu J, Descamps M, Dejou J, Koubi G, Hardouin P, Lemaitre J, Proust JP. The biodegradation mechanism of calcium phosphate biomaterials in bone. J Biomed Mater Res 2002;63:408-12. [45] Lu JX, Gallur A, Flautre B, Anselme K, Descamps M. Comparative study of tissue reactions to calcium phosphate ceramics among cancellous, cortical, and medullar bone sites in rabbits. J Biomed Mater Res 1998;42:357-67. [46] Doi Y, Shibutani T, Moriwaki Y, Kajimoto T, Iwayama Y. Sintered carbonate apatites as bioresorbable bone substitutes. J Biomed Mater Res 1998;39:603-10. [47] Liu Y, Cooper PR, Barralet JE, Shelton RM. Influence of calcium phosphate crystal assemblies on the proliferation and osteogenic gene expression of rat bone marrow stromal cells. Biomaterials 2007;28:1393-403. [48] Shelton RM, Liu Y, Cooper PR, Gbureck U, German MJ, Barralet JE. Bone marrow cell gene expression and tissue construct assembly using octacalcium phosphate microscaffolds. Biomaterials 2006;27:2874-81.

In: Environmental Biodegradation Research Focus Editor: B. Y. Wang, pp. 273-287

ISBN: 978-1-60021-904-7 © 2007 Nova Science Publishers, Inc.

Chapter 11

BIODEGRADATION OF PHENOL AND RESORCINOL BY A HALOTOLERANT PENICILLIUM Ana Lúcia Leitão∗ Grupo de Ecologia da Hidrosfera, Faculdade de Ciências e Tecnologia, Universidade Nova de Lisboa, Caparica, Portugal Unidade de Biotecnologia Ambiental, Quinta da Torre, Caparica, Portugal

ABSTRACT Many industries are known to generate wastewater enriched in phenolic compounds. These include petrochemicals, basic organic chemical manufacture, coal refining, pharmaceutical and tanning. Consequently, these compounds are commonly encountered in industrial effluents and surface water. Due to its high toxicity as shown by ecotoxicological studies, several methods have been reported for the removal of these pollutants from wastewater. Additionally to this toxicity problem some of these industrial effluents are likely to generate highly saline wastewaters. The discharge of such wastewaters containing at the same time phenol and phenolic compounds and high salinity without prior treatment is known to negatively affect the aquatic life, agriculture and potable water. Biological treatment with halotolerant/halophilic microorganisms is considered advantageous over the other physical and chemical methods as it leads to complete mineralization of phenolic compounds, is one of the safest, least costly and most socially acceptable. Halotolerant microorganisms are well known for their great versatility to remove pollutants, under saline and non saline conditions. Penicillium chrysogenum is an economically important ascomycete used as producer of penicillin. However, little attention has been paid to the ability of this microorganism to transform or metabolize compounds that are pollutants. This article presents a different approach to a classic problem. It employs an individual test on marine organism of trophic level 2 to

∗

Send correspondance to: Ana Lúcia Leitão; Grupo de Ecologia da Hidrosfera, Faculdade de Ciências e Tecnologia, Universidade Nova de Lisboa, Quinta da torre, 2829-516 Caparica, Portugal. Phone: +351-212948543, Fax : +351-21-2948543, E-mail : [email protected]

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Ana Lúcia Leitão validate bioremediation process by halotolerant fungus. A biodegradation process is an effective method of remediation if the toxicity of the system decreases. The purpose of this study was to compare cell growth and biodegradation of single phenol and resorcinol at high initial substrate levels by halotolerant strain, P. chrysogenum CLONA2, under saline and non saline conditions, adapted with either phenol or resorcinol, in a batch system, and to study the inhibitory and enhanced effect during the biodegradation of phenol and resorcinol. Single and binary substrate experiments were performed. HPLC analysis shows that halotolerant strain, Penicillium chrysogenum CLONA2, degraded up to 300 mg/l of both xenobiotics compounds in mineral salts medium with 58.5 g/l of sodium chloride. When phenol and resorcinol were together in low concentrations (< 15 mg/l and < 30 mg/l, respectively), phenol enhanced resorcinol degradation. P. chrysogenum CLONA2 metabolized phenol faster than resorcinol when present as the sole carbon source. The acute toxicity of phenol and resorcinol, individually and in combination, to larvae of the Artemia franciscana has been verified after and before bioremediation process with P. chrysogenum CLONA2. Resorcinol was more toxic than phenol. Our findings indicate that mixtures of resorcinol and phenol had an effect more toxic in A. franciscana than individual phenolic compounds.

Keywords: Resorcinol; Phenol; Halotolerant strain; Detoxification; Bioremediation

INTRODUCTION Phenol and resorcinol are aromatic molecules containing hydroxyl groups attached to the benzene ring structure. The origin of phenolic compounds in the environment is both natural as well as xenobiotic. Naturally aromatic compounds are widely distributed in plants, where these occur in the form of alkaloids, coumarins, flavonoids, terpenes, tannins and lignins. Phenols are also found in marine systems. Xenobiotic sources are industrial wastes derived from fossil fuel extraction as well from beneficiation and chemical manufacturing processes. Soil and aqueous environment are natural and preferential sinks for contamination, and their pollution represents an important concern for human and environmental health (Gianfreda and Bollag, 2002). Therefore, as phenol and resorcinol are chemicals of widespread use are commonly encountered in wastewater, surface water and in subsurface environments. Studies revealed that these phenolic compounds and their degradation products are one of the most aquatic pollutants, as important phenol was classified as “toxic” and resorcinol as “very toxic” (Kahru et al., 2000). Phenol is regarded as a priority contaminant by the U.S. Environmental Protection Agency (Keith and Telliard, 1979). The U.S. Environmental Protection Agency has set a limit of 600 μg/l as a 24-hour average to protect freshwater aquatic life, not exceeding 3.4 mg/l, and a drinking water limit of 1 μg/l (USEPA, 1979). In order to reduce the environmental load of harmful phenolic compounds, attention has to be focused on reducing the toxicity of wastewater by eliminating or decreasing the discharges of these toxic substances. In spite of, surface waters are frequently affected by a combination of toxicants operating in a variable environment. Usually, the interaction between two or more toxic chemicals is additive, however synergetic and antagonistic interactions are also been observed. So, if the

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toxicity of individual phenolic compounds in mixtures is additive, the chemical concentrations could be translated into toxicity data. The toxicity of phenol has been studied; however there are very little data available on the toxicity of resorcinol (Kahru et al., 2000) Various treatment methods have been proposed for phenolic compounds recovery or destruction. These methods include solvent extraction (Patterson, 1975), membrane separation (Chung et al., 1998), activated carbon adsorption (Nakhla et al., 1990), sonication (Mahamuni & Pandit, 2006), Fenton oxidation (Bittkau et al., 2004), electrochemical oxidation (Robson et al., 2007), ozonation (Azevedo et al., 2006) and biodegradation (Gaal & Neujahr, 1979; Godbole & Chakrabarti, 1991; Gunther at al., 1995; Bastos et al., 2000a; Margesin et al., 2004). In spite of some of these methods showed been effective in detoxification of phenols, most of them are not only expensive but also lead to the formation of secondary toxic materials or lower mineralization or severe operating conditions, and applicability to a limited phenol concentration range. Recently, a study proposes for mineralization of benzene revealed generates resorcinol and phenol not only as byproducts but also as end products. Indeed, Robson et al. (2007) showed that electrochemical oxidation over boron-doped diamond electrode in acidic medium was an efficient method in complete benzene oxidation. However, the complete mineralization of benzene yielded resorcinol and phenol that were not removed. Yao et al. (2006) employed a hydrogen peroxide as oxidizer and the enzyme from Serratia marcescens AB90027 to remove phenolic compounds. In this case, the degradation phenol and resorcinol were limited, in spite of some phenolic compounds were completely degraded by this method. Ozonation, in spite of to be a complex and expensive treatment, is the traditionally physicochemical method applied in effluents with phenol. However, new data showed that ozonation is an effective technology in low salinity media (2 g/l of sodium chloride), but at higher salinity levels, 50 g/l of sodium chloride, highly toxic compounds were produced (Azevedo et al., 2006). The salinity in most industrial effluents is a selective factor, and could be more toxic or stressful than phenolic compounds. Salinification, identified by European Union as a problem, is produced mainly as a result of anthropogenic activity. Due to its negative impact, the European Union Directive 2000/60/EC establishing a framework for Community action in the field of water policy in order to prevent saline pollution (European Union, 2000). Saline effluents are frequently treated by physicochemical methods, since it is known that salt decrease biological treatment efficiency as strongly inhibited microorganism. However, physicochemical methods are known to be energy consuming and the cost of the processes are high (Lefebvre and Moletta, 2006). Therefore, biological treatment of saline effluents is an alternative to be considered with microorganisms adapted to high ionic strength environment. A few species of bacteria are known to oxidize phenol under high salinity conditions (Hill et al., 1996; Chung et al., 1998; Bastos et al., 2000a; 2000b). However, there is little information about the degradative potential under saline environmental of halotolerant fungi. It is known that structurally-related compounds have an effect on its metabolic rates in a multi-substrate environment (Zollinger, 1966; Godbole and Chakrabarti, 1991; Latkar and Chakrabarti, 1994). Godbole and Chakrabarti (1991) reported biodegradation of resorcinol, catechol and phenol in mono and binary substrates matrices in upflow anoxic fixed film-fixed bed reactors. In this study under resorcinol acclimated reactor a binary mixed feed of phenol and catechol negatively interacted with one another. However, catechol and phenol acclimated reactors, showed in binary mixed feeds good substrate utilization efficiencies. Latkar and Chakrabarti (1994) demonstrated anaerobic degradation of resorcinol, catechol

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and hydroquinone in mono and binary substrate systems in upflow anaerobic fixed film-fixed bed reactors. In spite of several studies on phenol and resorcinol degradation efficiency under acclimated reactors, none of them have studied the efficiency of the acclimatization in aerobic conditions. A previous work has already investigated the degradation of phenol by Penicillium chrysogenum CLONA2 at high saline concentrations (Leitão et al., 2007). This strain could be an excellent candidate for bioremediation strategies in saline polluted environments. The aim of this study was to assess the efficiency of P. chrysogenum CLONA2 for removing phenol and resorcinol at high levels of salinity, 58.5 g/l of sodium chloride, without rendering apparently toxicity. This involved parallel testing of samples with A. franciscana tests. The influence of initial acclimation with phenol or resorcinol on biodegradability of these xenobiotics compounds was studied. The last step of this research was to compare the toxicity of monobasic and dibasic phenols.

2. MATERIAL AND METHODS 2.1 Strain One strain was isolated on complex medium as described (Leitão et al., 2007) from a salt mine of Clona in Portugal. The strain was able to grow in mineral medium with 58.5 g/l of sodium chloride and 200 mg/l of phenol as the sole carbon source. Phenol-degrading strain has been identified based on biochemical results by Leitão et al. (2007).

2.2 Growth medium The growth medium (Leitão et al., 2007) contained phenol, resorcinol or phenol and resorcinol as the carbon source, and mineral medium (MMFe) which contained the following components per litre: 1000 mg K2HPO4, 1000 mg (NH4)2SO4, 200 mg MgSO4.7H2O, 33 mg FeCl3.6H2O, 100 mg CaCl2, 58500 mg NaCl with pH adjusted to 5.6. All cultures in these experiments were incubated at 25±1 ºC in an INNOVA 4000 Incubator Shaker (New Brunswick Scientific, New Jersey, USA) operating at 160 rpm in the dark in order to avoid photo-destruction of phenol and resorcinol.

2.3 Inoculum development The strain was cultivated on nutrient agar (NA) plates (Difco Laboratories, Detroit, Michigan, USA) during 72 h of incubation at 25±1 ºC. Preculture of the microorganism was done from growth plate of strain in 100 ml of sterile complex medium (MC) which contained the following components per litre: 30.0 g glucose, 3.0 NaNO3, 0.5 g MgSO4.7H2O, 10 mg NH4Fe(SO4)2.12H2O, 1.0 g K2HPO4, 5.0 g yeast extract, 58.5 g NaCl with pH adjusted to 5.6. The preculture was incubated at 25±1 ºC for 3 days with shaking at 160 rpm.

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2.4 Effect of salt on resorcinol biodegradation Penicillium strain was cultivated at 0.0 and 5.9 % NaCl and 160 rpm in 250-ml Erlenmeyer flasks containing 50 ml mineral medium and 200 mg/l resorcinol. Three replicates were used per test salt concentration. Uninoculated control flasks (duplicates) were incubated and aerated in parallel as negative controls of the experiment. Growth and resorcinol concentration were monitored up to an incubation time of 96 h. Resorcinol concentration was quantified by HPLC as described in Liquid Chromatography section.

2.5 Acclimation For acclimation experiment, batch cultures were realized in mineral medium (MMFe) supplemented with 342 or 260 mg/l of phenol or resorcinol, respectively, under aerobic conditions. The optimization of culture parameter was done, in order to have a better degradation of phenolic compounds. After a period of acclimation of 76 hours, cultures were feed with resorcinol or phenol (300 mg/l). Cross feeding studies were conducted on all acclimated cultures using resorcinol or phenol as monosubstrate. The halotolerant strain was cultivated at 25 ºC and 160 rpm during 96 hours. Residual phenol and resorcinol concentrations were monitored at regular time intervals (24 h) by HPLC as described in Liquid Chromatography section.

2.6 Analytical methods Microbial dry biomass was estimated gravimetrically by centrifugation of 50 ml of each culture suspension at 4000 rpm for 10 min. Then the supernatant was discarded and the pellet was dried at 90ºC for 24 h and weighed (Gunther et al., 1995). Chromatographic analyses were performed in order to determine phenol and resorcinol degradation. The phenol and resorcinol concentrations were estimated by HPLC analysis (LaChrom HPLC Systeme, Merck), apparatus L-7100, equipped with a quaternary pump system, and a UV detector L7400. A personal computer equipped with HPLC System Manager Software for Windows NTTM was used to acquire and process chromatographic data. The separation was achieved with a LiChroCart 250-4 RP-18 endcapped (5 μm) column (Merck, Darmstadt, Germany). A mixture of deionized water/acetonitrile (70:30, v/v) under isocratic conditions was used as solvent and the flow rate was maintained at 1.0 ml/min. Analysis was carried out at 254 nm. Under these conditions, baseline separation for phenol and its intermediates could be obtained within 10 min. These compounds were identified by comparing their retention time with that of similarly treated external standards.

2.7 Acute toxicity test A 24-h LC50 bioassay was performed using instar II-III larvae of the brine shrimp A. franciscana according to the standardized ARTOXKIT MTM procedure. A reconstituted water

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of normal seawater salinity (35º/ºº) was prepared by adding reagent-grade chemicals to deionized water, and then used to prepare the hatching medium for the cysts and the dilution medium for the toxicant dilution series (ASTM Standard Guide E1440-91). A reference test with K2Cr2O7 was regularly performed for the control of the sensitivity of the test population.

2.8 Experimental design to study biodegradation and ecotoxicology assess of phenolic compounds Experimental studies were carried out with Erlenmeyer flasks, as batch reactors. A sample of 50 ml was taken in each 250 ml Erlenmeyer flask. Each sample contained the minimal medium with phenol, resorcinol or mixture (phenol and resorcinol) at different concentrations as the carbon source. The Erlenmeyer flasks were maintained at 25±1 ºC and shaking at 160 rpm, until phenolic concentrations to be less than 2 mg/l.

3. RESULTS

RESORCINOL CONCENTRATION (mg/l)

Batch cultures of P. chrysogenum CLONA2 were conducted in media containing initial resorcinol concentrations of 200 mg/l in the presence of 0 and 58.5 g/l of sodium chloride. Fig. 1 shows resorcinol concentration from representative experiments from each of the two conditions assayed. It can be seen that P. chrysogenum CLONA2 reduced resorcinol concentrations of 200 mg/l to below 5 mg/l in 96 h when 58.5 g/l of sodium chloride was added to the medium. When sodium chloride was not present in the medium resulted in a reduced rate of resorcinol removal. Resorcinol concentration was stable in the abiotic control. This result evidence the character halotolerant of P. chrysogenum CLONA2.

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Figure 1. Batch-degradation of 200 mg/l resorcinol by P. chrysogenum CLONA2 at 5.9% and 0% of sodium chloride.

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RESORCINOL CONCENTRATION (mg/l)

In order to assay the ability of P. chrysogenum CLONA2 to degrade resorcinol as the sole source of carbon and energy in 58.5 g/l of sodium chloride, an experiment was done with different initial concentrations of resorcinol, ranging from 50 to 300 mg/l (Fig. 2). The results showed that P. chrysogenum CLONA2 degraded at least 300 mg/l of resorcinol at high ionic strength. P. chrysogenum CLONA2 takes more time to degrade completely higher initial concentration of resorcinol. 300

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Figure 2. Degradation of resorcinol by P. chrysogenum CLONA2. Initial resorcinol concentration (mg/l), 50 (■), 75 (●), 100 (▲), 150 (○), 200 (♦), 250 (□), 300 (◊)

Figure 3. Resorcinol concentration (A) and growth profiles (B) during biodegradation of 200 mg/l of resorcinol

Fig. 3 shows the growth of P. chrysogenum CLONA2 and resorcinol degradation at initial concentration of 200 mg/l. At this initial concentration of resorcinol no lag phase was observed. Resorcinol degradation was well correlated with growth. Indeed, data indicate that there were two increments in biomass, the first one correspond to exponential growth phase the other one was observed at 69 h of culture. In both of the cases, biomass increments were

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coincident to resorcinol being consumed at faster rate. Since no abiotic loss of resorcinol was detected in controls, the measured of resorcinol disappearance could be attributed to biodegradation. Therefore, these results demonstrate that P. chrysogenum CLONA2 utilized resorcinol as carbon and energy source.

Figure 4. HPLC profiles of aerobic degradation of resorcinol (A, D), phenol (B, E) and phenol plus resorcinol (C, F) by P. chrysogenum CLONA2 at 28 h (A, B and C) and 48 h (D,E, and F) of culture. In the figures the phenol and resorcinol peaks are marked by arrows.

The degradation of phenolic compounds such as hydroquinone and catechol by P. chrysogenum CLONA2 at same conditions of assay than resorcinol has been observed (unpublished data). In fact, with the present study we demonstrated that P. chrysogenum CLONA2 could utilize the three compounds that have two hydroxyls as sole carbon and energy source. This ability is not often in microorganisms that have potential for bioremediation of phenolic compounds. The Arthrobacter sp. AG31 a cold tolerant

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microorganism and Pseudomonas putida DSM6414 a mesophilic one, are not able to utilize resorcinol for growth; on the other hand they are able to utilize phenol, catechol, hydroquinone between others aromatic hydrocarbons (Margesin et al., 2004). The position of hydroxyls has been reported as an important factor in the process of degradability of phenolic compounds. However, a few species of yeasts are known to grow on or metabolize at low concentrations phenol and resorcinol as sole sources of carbon. In this case it is not uncommon to found yeast that co-oxidize aromatic compounds (Cerniglia & Crow, 1981; Middelhoven, 1993). In this study, catechol and hydroquinone can be easily converted to benzoquinones, while degradation of resorcinol needs a process of hydroxylation or hydroxyl transference (Yao et al., 2006). The behavior of the strain in the presence of mixture phenolic compounds (phenol and resorcinol) was studied. Degradation efficiency of resorcinol, phenol and mixture phenolic compounds by P. chrysogenum CLONA2 was compared. Using batch cultivation in mineral medium with 58.5 g/l of sodium chloride and 50 mg/l of resorcinol, 50 mg/l of phenol was added. It was proved by chromatography analysis that P. chrysogenum CLONA2 metabolized at different rate phenol and resorcinol as the sole carbon source (Fig. 4). The degradation of resorcinol in the presence of phenol was initially delayed in the relation to the culture with only resorcinol. However, the rate of resorcinol degradation was higher in the culture with both monocyclic aromatic compounds (Fig. 5). Candida maltosa has been shown to tolerate resorcinol better than phenol, presumably because of differences in their chemical structure and in their respective degradation pathways via hydroxyquinol or protocatechuate (Fialová et al., 2004). The data presented here shows that P. chrysogenum CLONA2 metabolized phenol faster than resorcinol as the sole carbon source. 50

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Figure 5. Aerobic degradation of resorcinol (■), phenol (●), and resorcinol (▲) plus phenol (▼), by P. chrysogenum CLONA2

Experiments were performed in mineral medium with 58.5 g/l of sodium chloride at different concentrations of phenolic compounds (200 mg/l of phenol, 200 mg/l of resorcinol and mixture of 100 mg phenol/l and 100 mg resorcinol/l) to evaluate detoxification using

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Artemia salina (Fig. 6). According to the phenol study of Kahru et al. (2000), crustacean and/or photobacterial tests were most sensitive tests in the battery (Photobacterium phosphoreum, Microtox, Daphnia magna, Thamnocephalus platyurus, Tetrahymena thermophila, Selenastrum capricornutum) and exhibited complementary sensitivity patterns. However, Algaltoxkit and Protoxkit were least sensitive towards resorcinol (Kahru et al., 2000). As the present study was done in saline environment A. salina was the organism selected for this assay. According to Guerra (2001), A. salina could be a useful tool in assessing the toxicity of brackish effluents. 220

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100

CULTURE TIME (h)

Figure 6. The detoxification of phenolic compounds by P. chrysogenum CLONA2

Table 1. Effects of binary substrates feed in bioremediation process. Culture

Phenol

Resorcinol

Time (h)

(mg/l)

(mg/l)

Phenol (mg/l)

Resorcinol (mg/l)

0

201.6

201.7

107.6

105.9

24 48

185.8

174.3

71.9

83.2

78.9

94.9

0

20.5

72 96

0

28.3

0

2.2

0

6.7

0

2.0

Mixture

P. chrysogenum CLONA2 was able to degrade a mixture of 100 mg phenol/l and 100 mg resorcinol/l within 48-72 h of cultivation (Table 1 and Fig. 6). In the mixtures, phenol was completely degraded at 48 h of culture. However, a residual concentration of 2.0 mg/l of resorcinol was observed after 96 h of culture. High average levels of toxicity expressed by LC50 24-h was found in the samples from mixture cultures (phenol plus resorcinol) in the initial condition of the batch. Despite high percentages of removal of resorcinol, the samples taken from the batch at 72 h turned out to be toxic. However, at this time toxicity was not detected in the phenol batch and in the mixture of phenolic compounds. This absence of

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toxicity may be explained by the absence of phenolic compounds at concentrations that are toxic. There was a relation between the ability of phenolic compounds utilization and the toxicity of these compounds (Table 2). Phenol is least toxic compound than resorcinol. It is interesting to note that whereas the Microtox test performed with P. phosphoreum (Kaiser and Palabrica, 2000) displays the same high sensitivity to phenol as D. magna, T. platyurus and A. salina, the bacteria was ten times lower sensitive to resorcinol. On the contrary, LC50 value obtained with A. salina was between the intervals of values reported by Kahru et al. (2000). Table 2. Biological effects in P. phosphoreum, D. magna, T. platyurus and A. salina for the phenolic compounds during sampling periods

Phenol

Microtoxa 5 min EC50 (mg/l) 42.0

D. magna 48-h LC50 (mg/l) 8.3-

T. platyurusb 24-h LC50 (mg/l) 520

Resorcinol

264

0.2-

910

Compound

a b

A. salina 24-h LC50 (mg/l) 28.0 (27-32) 12.0 (6.5-14.0)

Kaiser and Palabrica (1991) Kahru et al (2000)

In order to gain an insight into the aerobic pathway of resorcinol, utilization of resorcinol and phenol was studied in acclimated cultures. Independent batch experiments with phenol and resorcinol as the only carbon source were carried out and the phenol and resorcinol concentrations profiles during consumption were monitored (Table 3). The initial phenolic concentrations was different, 342.0 and 260.0 mg/ for batch cultures with phenol and resorcinol, respectively. This experiment was design in order to have a residual phenolic concentration after 3 days of culture. Table 3. Reduction of phenol and resorcinol by P. chrysogenum CLONA2 Initial concentrations (mg/l)

Percent reduction after 24 h

Percent reduction after 48 h

Percent reduction after 72 h

Phenol

342

16.5

45.1

81.6

Resorcinol

260

36.8

78.5

95.4

Substrate

At this culture time, phenol and resorcinol biodegradation were 81.6 and 95.4, respectively. From the results reported in table 3 it seems that the ability of P. chrysogenum CLONA2 to use phenol (279 mg/l) is higher than to use resorcinol (248 mg/l), what are in agreement with the earlier experiments done. In spite of that it was found that at 24 h of culture resorcinol degradation rate was higher than phenol degradation rate. This result could

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be explain by the fact that at high concentrations of phenol the metabolic intermediates, from cleavage of aromatic ring, could be much more reactive and compete with phenol in the series of enzymatic reactions, resulting in inhibition on phenol consumption. Table 4 shows the results of feeding resorcinol and phenol in batch cultures of P. chrysogenum CLONA2 performed in media containing 58.5 g/l of sodium chloride with 300 mg/l of these phenolic compounds. Table 4. Biodegradation of phenol and resorcinol in phenol and resorcinol acclimated cultures

Batch Phenol

Feed Resorcinol Phenol

Percent Reduction after 24 h 71.4 55.0

Resorcinol

Resorcinol Phenol

35.6 18.3

Removal Reduction after 48 h

75.9 43.1

97.4 99.7

Reduction of resorcinol and phenol in resorcinol acclimated cultures was similar to phenolic reductions in batch cultures. While in phenol acclimated cultures, phenol and resorcinol could be utilized consistently to the extent of 99.7 and 97.4%, respectively, throughout 48 h of culture. Further, resorcinol degradation at 24 h of culture, even when present with residual phenol concentration, was improved in the phenol acclimated cultures; resorcinol reduction being 71.4%. Phenol degradation was 55 % after the first day of initiation of the phenol acclimated experiment, suggesting that under these culture conditions resorcinol is more easily use than phenol. These observations suggest that phenol seems to be a better inducer of the reductive pathway than resorcinol as is evidenced by resorcinol and phenol utilization efficiency of the phenol acclimated cultures. Several studies have shown that the first enzyme of the phenol degradation pathway (phenol hydroxylase) has broad substrate specificity (Neurajahr & Varga, 1978; Detmer & Massey, 1985). The phenol hydroxylase from Tricosporon cutaneum has been reported to catalyse the hydroxylation of a variety of substituted phenols. Neurajahr and Varga (1978) reported that cells grown on resorcinol contain enzymes that participate in the degradation of phenol and vice versa. Indeed we detected phenol hydroxylase activity when substrate was not only phenol but also resorcinol (data not shown). These results demonstrate the efficiency of P. chrysogenum CLONA2 for biodegradation processes at high saline concentrations. Bioremediation in high saline concentrations environment is based on the ability of halo-adapted microorganisms to degrade xenobiotics compounds under saline conditions. The environment is commonly polluted by several contaminants at the same time. For instance, the contamination of soils and water by fossil fuel refining process industries wastes, where phenol and resorcinol are predominant aromatics present is often accompanied by soil salinization. Therefore, phenol and resorcinol biodegradation is often a prerequisite for the treatment of mixed pollutants where they are present in relatively high concentrations.

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Phenolic compounds and salinity affect adversely the efficiency of wastewater treatment (Kargi and Dinçer, 1999). The use of salt-adapted microorganisms that are able of face high salinities and at the same time of removing pollutants that are present wastewater is recommended in the treatment of saline effluents, before salt removal by physicochemical methods (Lefebvre and Moletta, 2006). The evaluation of chemical and ecotoxicological data was useful for predicting the effect of the raw effluent on the treatment plant and the impact of the final treated effluent on the receiving water (Guerra, 2001).

4. CONCLUSION P. chrysogenum CLONA2 was able to utilize resorcinol as the sole carbon source over a wide range of resorcinol and total salt concentrations. Biodegradation of resorcinol is more efficient at 5.8 % (w/v) NaCl than in the absence of salt. The two phenolic compounds tested in this study have different LC50, and consequently exhibit different toxicity. Resorcinol was more toxic than phenol. There were differences in detoxification efficiencies of phenolic compounds. In the case of batch with phenol or mixture (phenol plus resorcinol) phenol was most rapidly detoxified. When phenol and resorcinol were together in low concentrations, phenol enhanced resorcinol degradation. The mixture of phenolic compounds had an effect more toxic in A. franciscana than individual phenolic compounds. Resorcinol pathway is better expressed in the phenol acclimated cultures than in the resorcinol acclimated cultures. Attend to all findings present in this study the possibility to use this halotolerant strain of P. chrysogenum for detoxification of resorcinol and phenol saline wastewater, must be considered.

REFERENCES Azevedo, E .B., Aquino de Neto , F. R., & Dezotti, M. (2006). Lumped kinetics and acute toxicity of intermediates in the ozonation of phenol in saline media. Journal of Hazardous Materials, 128, 182-191. Bastos, A. E., Moon, D. H., Rossi, A., Trevors, J. T., & Tsai, S. M. (2000a). Salt-tolerant phenol-degrading microorganisms isolated from Amazonian soil samples. Archives of Microbiology, 174, 346-352. Bastos, A. E., Tornisielo, V. L., Nozawa, S. R., Trevors, J. T., & Rossi, A. (2000b). Phenol metabolism by two microorganisms isolated from Amazonian forest soil samples. Journal Industrial Microbiology & Biotechnology, 24, 403-409. Bittkau, A., Geyer, R., Bhatt, M., & Schlosser, D. (2004). Enhancement of the biodegradability of aromatic groundwater contaminants. Toxicology, 205, 201-210. Cerniglia, C. E., & Crow, S. A. (1981). Metabolism of aromatic hydrocarbons by yeasts. Archives of Microbiology, 129, 9-13. Chung, T. S., Loh, K. C., & Tay, H.L. (1998). Development of polysulfone membranes for bacteria immobilization to remove phenol. Journal of Applied Polymer Science, 70, 2585-2594.

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Detmer, K., & Massey, V. (1985). Effect of substrate and pH on the oxidative half-reaction of phenol hydroxylase. The Journal of Biological Chemistry, 260, 5998-6005. European Union (2000). Directive 2000/60/EC of the European Parliament and the Council of 23 October 2000 establishing a framework for Community action in the field of water policy. In: Official Journal of the European Community L327, 1-73. Fialová, A., Boschke, E., & Bley, T. (2004). Rapid monitoring of the biodegradation of phenol-like compounds by the yeast Candida maltosa using BOD measurements. International Biodeterioration & Biodegradation, 54, 69-76. Gaal, A., & Neujahr, H. Y. (1979). Metabolism of phenol and resorcinol in Tricosporon cutaneum. Journal of Bacteriology, 137, 13-21. Gianfreda, L., & Bollag, J. (2002). Isolated enzymes for the transformation and detoxification of organic pollutant. In Burns, R., & Dick, R. (Eds.) Enzymes in the environment: activity, ecology, and application. 495-538. New York: Marcel Dekker. Godbole, A., & Chakrabarti, T. (1991). Biodegradation in upflow anoxic fixed film-fixed bed reactors of resorcinol, catechol and phenol in mono and binary substrates matrices. Water Research, 25, 1113-1120. Gunther, K., Schlosser, D., & Fritsche, W. (1995). Phenol and cresol metabolism in Bacillus pumilus isolated from contaminated groundwater. Journal of Basic Microbiology, 35, 8392. Guerra, R. (2001). Ecotoxicological and chemical evaluation of phenolic compounds in industrial effluents. Chemosphere, 44, 1737-1747. Hill, G. A., Milne, B. J., & Nawrocki, P. A. (1996). Cometabolic degradation of 4chlorophenol by Alcaligenes eutrophus. Applied Microbiology and Biotechnology, 46, 163-168. Kahru, A., Pollumaa, L., Reiman, R., Ratsep, A., Liiders, M., & Maloveryan, A. (2000). The toxicity and biodegradability of eight main phenolic compounds characteristic to the oilshale industry wastewaters: a test battery approach. Environmental Toxicology, 15, 431442. Kaiser, K. L. E., & Palabrica, V. S. (2000). Photobacterium phosphoreum toxicity data index. Water Pollutant Research Journal of Canada, 26, 361-431. Kargi, F., & Dinçer, A. R. (1999). Salt inhibition effects in biological treatment of saline wastewater in RBC. Journal of Environmental Engineering, 125, 966-971. Keith, L. H., & Telliard, W. A. (1979). Priority pollutants. I. A prespective view. Environmental Science and Tecnology, 13, 416-423. Latkar, M., & Chakrabarti, T. (1994). Resorcinol, catechol and hydroquinone biodegradation in mono and binary substrate matrices in upflow anaerobic fixed-film fixed-bed reactors. Water Research, 28, 599-607. Lefebvre, O., & Moletta, R. (2006). Treatment of organic pollution in industrial saline wastewater: A literature review. Water Research, 40, 3671-3682. Leitão, A. L., Duarte, M. P., & Oliveira, J. S. (2007). Degradation of phenol by a halotolerant strain Penicillium chrysogenum. International Biodeterioration and Biodegradation, 59, 220-225. Mahamuni, N. N., & Pandit, A. B. (2006). Effect of additives on ultrasonic degradation of phenol. Ultrasonics Sonochemistry, 13, 165-174.

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Margesin, R., Bergauer, P., & Gander, S. (2004). Degradation of phenol and toxicity of phenolic compounds: a comparison of cold-tolerant Arthrobacter sp. and mesophilic Pseudomonas putida. Extremophiles, 8, 201-207. Middelhoven, W. J. (1993). Catabolism of benzene compounds by ascomycetous and basidiomycetous yeasts and yeast-like fungi. Antonie Leeuwenhoek, 63, 125-144. Nakhla, G. F., Suidan, M. T., & Pfeffer, J. T. (1990). Control of anaerobic GAC reactors treating inhibitory wastewaters. Journal of Water Pollution Control Federation, 62, 6572. Oliveira, R. T. S., Salazar-Banda, G. R., Santos, M. C., Calegaro, M. L., Miwa, D. W., Machado, S. A. S., & Avaca, L. A. (2007). Electrochemical oxidation of benzene on boron-doped diamond electrodes. Chemosphere, 66, 2152-2158. Patterson, J.W. (1975). Wastewater Treatment Technologies. Ann Arbor, Michigan, Ann Arbor Science Publishers. USEPA (1979). Phenol. In Ambient Water Criteria. PB 296, 787. Office of Water Planning and Standards, U.S. Environmental Protection Agency, Washington, DC. Zollinger, E. S. (1966). Effects of inhibition and repression on the utilization of substrates by heterogeneous bacterial communities. Applied Microbiology, 14, 654-664. Yao, R. S., Sun, M., Wang, C. L., & Deng, S. S. (2006). Degradation of phenolic compounds with hydrogen peroxide catalyzed by enzyme from Serratia marcescens AB 90027. Water Research, 40, 3091-3098.

In: Environmental Biodegradation Research Focus Editor: B. Y. Wang, pp. 289-302

ISBN: 978-1-60021-904-7 © 2007 Nova Science Publishers, Inc.

Chapter 12

KINETICS AND METABOLIC PATHWAY OF MELATONIN BIODEGRADATION BY A BACTERIUM ISOLATED FROM THE MANGROVE SEDIMENT Xiang-Rong Xu 1, 2, ∗, Hua-Bin Li 1, Ji-Dong Gu 1 and Xiao-Yan Li 2 1

Department of Ecology & Biodiversity, The University of Hong Kong, Pokfulam Road, Hong Kong 2 Department of Civil Engineering, The University of Hong Kong, Pokfulam Road, Hong Kong

ABSTRACT Melatonin (MLT) is a hormone produced primarily by the pineal gland. It can be found in animals and humans as well as a number of bacteria, fungi and plants. Since it has been widely used as the healthcare product in the world market, MLT entered the various environmental compartments by different ways. Recent research found that MLT may suppress the production of testosterone, decrease semen quality, and affect sexual activity and reproduction of animal and human. The fate of melatonin in environment has attracted increasing public and scientific concerns in recent years. In this chapter, biodegradability of MLT was studied for the first time. A strain that can efficiently degraded melatonin was isolated from the mangrove sediment. This strain, a gramnegative bacterium identified as Shewanella putrefaciens, can grow on melatonin as sole sources of carbon and energy under aerobic conditions. The growth was greatly enhanced by the addition of a small amount of yeast extract. Effects of melatonin concentration, pH, temperature and salinity on MLT biodegradation were studied, respectively. The experimental results showed that 50 mg l-1 melatonin could be degraded within 2 d under the optimal condition (pH 7.0, salinity 15‰ and temperature at 37 °C). The process of MLT biodegradation was monitored by high-performance liquid chromatography with ultra-violet detection. The biodegradation of melatonin could be fitted to a first-order kinetic model. The major metabolites of melatonin biodegradation were identified by

290

Xiang-Rong Xu, Hua-Bin Li, Ji-Dong Gu and Xiao-Yan Li high-performance liquid chromatography and gas chromatography-mass spectrometry, and a preliminary metabolic pathway of melatonin was proposed. The results obtained are helpful to understand environmental behavior of MLT, and also could be used for the bioremediation of MLT-contaminated site, such as the wetland of the Mai Po Natural Reserve in Hong Kong.

Keywords: Melatonin; Biodegradation; Kinetics; Metabolic pathway

1. INTRODUCTION Melatonin (MLT), N-acetyl-5-methoxytryptamine, is an indole hormone produced primarily by the pineal gland and retina [Jobling et al., 1998; Reiter, 1991]. It can be found in animals and humans as well as a number of bacteria, fungi and plants [Kolar & Machackova, 2005; Tan et al., 2003]. Recent research found that MLT may suppress the production of testosterone, decrease semen quality, and affect sexual activity and reproduction of animal and human [Luboshitzky et al., 2002; Tanyildizi et al., 2006; Yamada et al., 1992; Yilmaz et al., 2000]. Since it has been widely used as the healthcare product in the world market, MLT entered the various environmental compartments by different ways. The fate of melatonin in environment has attracted increasing public and scientific concerns in recent years. In this chapter, biodegradability of MLT was studied for the first time. The Mai Po Natural Reserve is the largest remaining wetland in Hong Kong and plays a very important role in supporting a wide range of wild life. Its mangrove ecosystem, important inter-tidal estuarine wetland along the coastlines of tropical and sub-tropical regions, is closely associated with human activities and is subject to different organic pollutants (including melatonin) contamination [Xu et al., 2005; Xu et al., 2007]. The mangrove sediment might contain melatonin-degrading bacteria. In this chapter, a pure culture of bacteria capable of using melatonin as the sole carbon and energy source was isolated from the mangrove sediment in Hong Kong. Effects of melatonin concentration, yeast extract, hydrogen peroxide, pH value, temperature and salinity on MLT biodegradation were studied, respectively. The process of MLT biodegradation was monitored by highperformance liquid chromatography with ultra-violet detection. The major metabolites of melatonin biodegradation were identified by high-performance liquid chromatography and gas chromatography-mass spectrometry. The results obtained are helpful to understand environmental behavior of MLT, and also could be used for the bioremediation of MLTcontaminated site, such as the wetland of the Mai Po Natural Reserve in Hong Kong.

∗

Corresponding author: Tel.: 00852-2859-1970; Fax: 00852-2559-5337; E-mail: [email protected].

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291

2. MATERIALS AND METHODS 2.1. Chemicals All reagents were of analytical-reagent grade and purified water by Milli-Q system was used throughout the experiments. Melatonin (99.5%) was purchased from Aldrich (St Louis, MO, USA). HPLC-grade methanol was purchased from TEDIA (Ohio, USA). Other reagents were purchased from Sigma.

2.2. Enrichment culture The initial enrichment culture was established by inoculating a 250 ml Erlenmeyer flask containing 100 ml mineral salt medium (MSM) supplemented with melatonin (20 mg l−1) as the sole carbon and energy source with 5 g fresh sediment from Mai Po Nature Reserve in Hong Kong. The MSM comprised (mg l−1): (NH4)2SO4 1000; KH2PO4 800; K2HPO4 200; MgSO4·7H2O 500; FeSO4 10; CaCl2 50; NiSO4 32; Na2BO7·H2O 7.2; (NH4)6Mo7O24·H2O 14.4; ZnCl2 23; CoCl2·H2O 21; CuCl2·2H2O 10 and MnCl2·4H2O 30. The pH of the culture medium was adjusted with HCl or NaOH to 7.0 ± 0.1 or as otherwise specified. The flasks were incubated in the dark in an INNOVA 4340 Incubator Shaker (New Brunswick Scientific, New Jersey, USA) operating at 150 rpm and 25.0 ± 0.5 °C. The melatonindegrading cultures were obtained through enrichment transfer at approximately 1-week intervals on the basis of depletion of melatonin, by transferring 1.0 ml of the active culture to a new Erlenmeyer flask containing 100 ml of freshly made MSM with gradually increasing concentrations of melatonin. The melatonin-degrading enrichment cultures were transferred more than 10 times prior to the isolation of bacteria from the enrichment cultures.

2.3. Isolation and characterization of microorganism Bacteria in the enrichment culture showing ability in degrading melatonin were diluted in MSM before plating on the nutrient agar (NA) plates (Difco Lab., Detroit, Michigan). After 48 h of incubation at 25 °C, a number of well-separated, individual colonies of different colony morphological types appeared and were further streaked onto fresh NA plates (Difco Lab., Detroit, Michigan) to purify the cultured organisms. Pure cultures were used subsequently for Gram staining and then identification using API 20 NE Multi-test System (bioMerieux, Marcy 1’Etoile, France) as described elsewhere (Wang et al., 2004).

2.4. Biodegradation of melatonin by the isolated strain All flasks were incubated at 25 °C on a rotary shaker operated at 150 rpm in the dark or otherwise specified. Experiments on melatonin degradation were conducted in 250 ml Erlenmeyer flasks with 50 ml MSM. The following factors were then adjusted or added in order to study their effects on MLT degradation: pH (5, 6, 7, 8 and 9); temperature (20, 25,

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Xiang-Rong Xu, Hua-Bin Li, Ji-Dong Gu and Xiao-Yan Li

30, 37 and 40 °C); salinity (0, 10, 15, 20, 30 and 35 ‰); yeast extract (0.5 mg l-1); hydrogen peroxide (1.0 mg l-1). All tests were conducted in triplicate. Sterile controls were prepared by autoclaving before introduction of melatonin, which passed through a membrane filter 0.2-μm pore size (Pall Gelman Laboratory, Ann Arbor, Michigan).

2.5. Solid-phase extraction procedure A Waters Sep-Pak C18 cartridge (500 mg) was conditioned by 2 ml of methanol, and then 5 ml of water [Xu et al., 2006]. The sample (10 ml of culture medium) was acidified to pH 2 with 0.1 N HCl, and then passed through the cartridge at a flow-rate of 2 ml min-1. The analytes retained on the solid-phase extraction (SPE) cartridge was eluted with methanol (1 ml × 2), and the eluate was purged to dryness with pure nitrogen gas. Finally, the residue was dissolved in 0.5 ml of methanol prior to determination by high-performance liquid chromatography with ultra-violet detection or gas chromatography-mass spectrometry (GCMS). The recoveries of melatonin after SPE were approximately 95%. All tests were conducted in triplicate.

2.6. Determination of melatonin and identification of its metabolites The Agilent 1100 series HPLC consisting of a G1322A degasser, a G1311A QuatPump, a G1316A COLCOM and a G1315B diode array detector (DAD, Agilent, USA) set at 225 nm wavelength, was used for the detection of melatonin concentration. A personal computer equipped with a HP ChemStation (HP, USA) was used to acquire and process chromatographic data. An Agilent Zorbax Eclipse XDB-C8 column (150 × 4.6 mm, particle size 5 μm) was used for separation. The mobile phase was a mixture of methanol and water (40:60, v/v) and the flow rate was set at 0.8 ml min−1. Under these chromatographic conditions, baseline separation can be obtained within 20 min for melatonin and its metabolites. The peak area was integrated and used as the analytical signal for quantification. All compounds studied were quantified using external standards. Identification of degradation metabolites of melatonin was performed using a Hewlett Packard 6890 gas chromatograph (Hewlett Packard, USA) equipped with an Agilent 5973 mass selective detector (Agilent, USA). The column used was a HP-624 silicone-coated, fused-silica capillary column (25.0 m × 0.2 mm i.d., 1.12 μm film thickness). The injector and detector temperatures were 290 and 310 °C respectively. The initial oven temperature was 120 °C. Then the temperature was ramped to 280 °C at 10 °C min−1, and held for 9 min. The injection volume is 1 μl and the carrier gas was helium (1.0 ml min−1). A mass range of 30−550 amu was scanned in all electron ionization mass spectroscopy studies where the electron energy was 70 eV. Instrumental library searches, comparison with available standard compounds, and mass fragmentation pattern were used to identify the suspected metabolites.

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3. RESULTS AND DISCUSSION 3.1. Isolation and characterization of a melatonin-degrading strain A melatonin-degrading microbe was isolated by enrichment shaking culture at 25 °C. The strain was purified by successive streak transfers on agar-plate medium and maintained as slant cultures on nutrient broth agar. The experiment showed that this melatonin degrader was capable of using melatonin as the sole source of carbon and energy. It was identified as Shewanella putrefaciens with 98.6% similarity by API 20NE system preliminarily. It was a gram-negative, non-motile, rod-shaped, facultative aerobic bacterium of 0.5~0.7 μm in diameter and 0.7~2.2 μm in length.

3.2. Effect of the initial melatonin concentration By measuring the changes of optical density at 600 nm, the melatonin concentration ranged from 25 to 300 mg l−1 was determined. Between 25 and 150 mg l−1, microbial growth increased with melatonin concentration, but at 200 mg l−1 of melatonin, the growth rate was lower and at 300 mg l−1 growth was totally inhibited (Figure 1). Thereafter, all experiments were carried out using < 200 mg l−1 of melatonin.

Absorbance 600 nm

0.40

0.30

0.20

0.10

0.00 0

12

24

36

48

60

72

84

96

Time (h) Figure 1. Effect of melatonin concentrations on microbial growth measured as optical density at 600 nm. (□: 25 mg l−1; ■: 50 mg l−1; Δ: 100 mg l−1; ▲: 150 mg l−1; ◊: 200 mg l−1; ♦: 300 mg l−1. Values are means of three different assays measured in duplicate (n = 6); error bars correspond to one standard deviation.

Biodegradation rate of organic compounds by microorganisms is often described by the equation as follows: γ = γm c/(k + c)

(1)

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Xiang-Rong Xu, Hua-Bin Li, Ji-Dong Gu and Xiao-Yan Li

where γ is biodegradation rate, γm is maximum specific biodegradation rate, c is the substrate concentration and k is half-saturation constant. If c<
(2)

Above Eq. (2) is a typical first-order model. Assuming k1 = (γm/k) and integrating Eq. (2), the following relation of substrate concentration to time can be obtained: ln c = a – k1t

(3)

If c>>k, another simplified equation can be got from Eq (1): γ = γm

(4)

This biodegradation is zero-order and the rate constant k0 = γm. Thus the relationship between substrate concentration and time is: (5) C = b + k0t The biodegradation of melatonin by microbial cells was investigated with various initial concentrations of melatonin. The results were shown in Figure 2.

-1

[Melatonin] (mgl )

250 200 150 100 50 0 0

24

48

72

96

120

144

Time (h) Figure 2. Effect of initial concentration of melatonin on degradation by Shewanella putrefaciens at 30°C and pH 7.0. Initial concentrations of melatonin used (mg l−1) were: 25(□), 50(■), 75( ), 100(◊); 150 (♦).

It can be seen that when the initial concentrations of melatonin varied between 25 and 150 mg l−1, it could be completely degraded by the cells of Shewanella putrefaciens within 144 h. According to above kinetic analysis, concentration (C) ~ time (t) curve (Figure 2) and ln C~t curve (Figure 3) were plotted to fit the data. It was found that biodegradation kinetics

Kinetics and Metabolic Pathway of Melatonin Biodegradation…

295

of melatonin by bacteria can be described well with first-order model when initial melatonin concentration was in the range of 25 to 150 mg l−1. The kinetic results of melatonin biodegradation were summarized in Table 1. Table 1 indicated that melatonin was utilized by the microbial cells at a constant rate when the initial concentration of melatonin was in the range of 25 to 150 mg l−1. The relationship between the different initial substrate concentrations and the initial degradation rates was linear (r2 >0.95).

ln [Melatonin]

6

4

2

0

-2 0

24

48

72

96

120

144

Time (h) Figure 3. Biodegradation of melatonin and first-order reaction model prediction by Shewanella putrefaciens at 30°C and pH 7.0. Initial concentrations of melatonin used (mg l−1) were: 25(□), 50(■), 75( ), 100(◊); 150 (♦).

Table 1. Kinetic equations of melatonin biodegradation by Shewanella putrefaciens Initial Concentration (mg l−1)

Kinetic equations

Rate constants

r2

(h−1) 25

ln C = 3.37 – 0.030t

0.030

0.95

50

ln C = 4.09 – 0.031t

0.031

0.98

75

ln C = 4.49 – 0.031t

0.031

0.99

100

ln C = 4.69 – 0.030t

0.030

0.97

150

ln C = 5.25 – 0.031t

0.031

0.96

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Xiang-Rong Xu, Hua-Bin Li, Ji-Dong Gu and Xiao-Yan Li

3.3. Effects of yeast extraction and hydrogen peroxide on melatonin biodegradation Data on the effects of the addition of yeast extract, or hydrogen peroxide on MLT degradation are presented in Table 2. The MLT degradation rate constant calculated was 0.038 h−1 and the half-life was 18.24 hrs in the inoculated control. As the results show, MLT degradation was enhanced by the addition of yeast extract. The results revealed relatively good growth and degradation capacity in the presence of yeast extract. This is consistent with other studies [Chang et al., 2005]. MLT degradation was inhibited by the addition of hydrogen peroxide, the toxicity of which decreased the activity of the microorganisms. Table 2. Effects of yeast extract and hydrogen peroxide on melatonin degradation rate constants (k1) and half-lives (t1/2). k1 (h−1)

t1/2 (h)

ra

Inoculated controlb

0.038

18.24

0.93

Yeast extract

0.043

16.12

0.95

Hydrogen peroxide

0.031

22.36

0.92

Treatment

Each figure represents the mean of three measurements; in all cases, standard deviation was less than 10%. All inoculated control and treatment figures were significantly different at p < 0.05. a r = correlation coefficient. b Inoculated control: 30 °C, pH 7.0, and 50 mg l−1 of melatonin.

-1

Rate constant (h )

0.080 0.060 0.040 0.020 0.000 5

6

7

8

pH Figure 4. Effect of pH on melatonin biodegradation at 30° C. [melatonin]0 = 50 mg l−1.

9

Kinetics and Metabolic Pathway of Melatonin Biodegradation…

297

3.4. Effect of pH on melatonin biodegradation The hydrogen ion concentration in the culture medium could influence the bacterial growth and activity of enzymes. The relationship between the degradation rate constants and pH for Shewanella putrefaciens is shown in Figure 4. The rate constants of melatonin degradation increased gradually when pH value of the culture was increased from 5.0 to 7.0. A highest rate constant was achieved for Shewanella putrefaciens at pH 7.0. When the pH was greater than 7.0, the rate constants decreased. Thus, the pH value for melatonin degradation by Shewanella putrefaciens was chosen as 7.0 in subsequent experiments.

3.5. Effect of temperature on melatonin biodegradation When the effect of temperature on MLT degradation by Shewanella putrefaciens was assessed, the rate constants increased with the increase of temperature between 20 °C and 37 °C (Figure 5). Higher temperature resulted in the lowering of the degradation rate constant; the optimum was 37 °C, at which temperature the degradative enzyme reached the highest activity for the temperature tested in this study.

-1

Rate constant (h )

0.080 0.060 0.040 0.020 0.000 20

25

30

35

40

o

Temperature ( C) Figure 5. Effect of temperature on melatonin biodegradation at pH 7.0. [melatonin]0 = 50 mg l−1.

3.6. Effect of salinity on melatonin biodegradation Effect of salinity on MLT biodegradation by Shewanella putrefaciens was assessed and the results are presented in Figure 6. In a simulated freshwater culture, the rate constant was 0.0309 h−1. With the increase of salinity from 0 to 15‰, the rate constant increased greatly. The highest MLT degradation was found at 15‰. Higher salinity than 15‰ resulted in a decrease of degradation rate mainly due to the inhibition to the growth of Shewanella putrefaciens.

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Xiang-Rong Xu, Hua-Bin Li, Ji-Dong Gu and Xiao-Yan Li

Based on the above optimal conditions obtained, we carried out the melatonin biodegradation by Shewanella putrefaciens. The experimental results indicated that, at pH 7.0, temperature 37 °C and salinity 15‰, 50 mg l-1 melatonin could be degraded by Shewanella putrefaciens within 2 d.

-1

Rate constants (h )

0.080

0.060

0.040

0.020 0

5

10

15

20

25

30

35

Salinity (‰ ) Figure 6. Effect of salinity on melatonin biodegradation at pH 7.0 and at 30° C. [melatonin]0 = 50 mg l−1.

3.7. Metabolic pathway of melatonin biodegradation by Shewanella putrefaciens To explore the metabolic pathways of melatonin by Shewanella putrefaciens, the culture medium was extracted using solid-phase extraction, and the extract was separated by highperformance liquid chromatography (HPLC) with ultraviolet detection. Besides the parent compound melatonin, two major metabolites were identified by comparison with standard compounds. Other metabolites with very small peaks could not be isolated for further characterization. Melatonin degradation products were also analyzed by gas chromatographymass spectrometry (GC-MS), and three main compounds were identified by comparing the mass spectrum at particular retention time with the published mass spectra database. The result obtained by GC-MS was consistent with that by HPLC. The separation and identification methods/procedures of melatonin metabolites were very similar to those reported in the literature [Xu et al., 2005, Xu et al., 2006]. Based on the above results, combining with the literature [Gu & Berry, 1991; Gu & Berry, 1992; Kaiser et al., 1996; Yin et al., 2006], a biochemical pathway for metabolism of melatonin by Shewanella putrefaciens might be proposed (Figure 7).

Kinetics and Metabolic Pathway of Melatonin Biodegradation…

CH3O

CH2

CH2

NHCOCH3

CH2

CH2

NHCOCH3

299

N H

CH3O N

OH

H

CH3O

CH2 N

CH2

NHCOCH3

CH2

NHCOCH3

O

H

CH3O

CH2 NH2

COOH

CO2 + H2O Figure 7. A proposed pathway for melatonin biodegradation by Shewanella

putrefaciens.

In the literature, there is no information on the biodegradation of melatonin. In addition, information on degradation, transformation and the fate of indolic class chemicals is extremely limited considering the vast amounts of information available on chlorinated

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aromatics and polyaromatics [Berry et al., 1987; Gibson, 1984; Kaiser et al., 1996; Young and Cerniglia, 1995]. Only one chemical in the indolic class has been investigated to a limited extent, the chemical indole was investigated under sulfate-reducing condition [Bak and Widdel, 1986], denitrifying condition [Madsen and Bollag, 1989] and methanogenic conditions [Gu and Berry, 1991; Gu and Berry, 1992]. Among these studies, only one sulfatereducing bacterium was isolated and identified as Desulfobacterium indolicum by Bak and Widdel [1986]. Other investigations using indole as a test chemical emphasized the degradability of the chemical and the pathway of biodegradation, and degradation pathway was partially established [Berry et al., 1987; Gu and Berry, 1991; Madsen and Bollag, 1989]. Further study also revealed that incubation temperature and the amount of sediment or sludge inoculum used also influenced the rate of degradation [Madsen et al., 1988]. Gu et al. [2002] found that two intermediates identified as oxindole and isatin (indole-2,3-dione) had been produced during indole biodegradation under both methanogenic and sulfate-reducing condition. The results suggested that both methanogenic and sulfate-reducing bacteria use an identical degradation pathway. Degradation processes followed two steps of oxidation accomplished by hydroxylation and then dehydrogenation at 2- and then 3-position sequentially prior to the cleavage of the pyrrole ring between 2- and 3-positions. However, none of 1-methylindole or 2-methylindole was degraded under any conditions. 3Methylindole (3-methyl-1H-indole, skatole) was transformed under methanogenic conditions and mineralized only under sulfate-reducing conditions. It is clear that methyl substitution on 1- or 2-position inhibits the initial attack by hydroxylation enzymes making them more persistent in the environment and posing longer toxic impact. Yin et al. [2006] investigated the biodegradation of 1-methylindole (1MI) and 3-methylindole (3MI) using enrichment cultures with mangrove sediment obtained from Mai Po of Hong Kong. A pure culture Pseudomonas aeruginosa Gs could use 1MI and 3 MI as the sole source of carbon and energy. Oshima et al. [1965] reported that indole was oxidized by Pseudomonas indoloxidans to indoxyl, which then oxidized spontaneously to indigotin. However, Ps. indoloxidans could not utilize indole either as a carbon or a nitrogen source [Oshima et al., 1965]. No blue pigment was observed when cultures in MSM containing indole were incubated. However, when a small quantity of yeast extract (0.5 g l−1) was added, the culture medium turned blue gradually indicating the formation of indigo, so showing that with yeast extract indole could be metabolized to indigo by Ps. aeruginosa Gs. Claus and Kutzner [1983] found that Alcaligenes sp. (strain In3) could degrade indole as sole carbon source under aerobic conditions and isatin and gentisate were the major intermediates, but strain In3 could not degrade skatole (3-methylindole). Otherwise, in water–organic solvent system of two phases, indole was degraded by Pseudomonas sp. ST-200, first being oxygenated at 2- and 3positions to form cis-indole-2,3-dihydrodiol, then oxidized to form isatin, and subsequently cleaved at the C–N bond between 1- and 2-positions to form isatinic acid, which was not assimilated by ST-200 as a growth substrate [Doukyu and Aono, 1997]. However, isatin was not observed in the culture medium of the current study and unidentified intermediates of indole were also degraded without accumulation. In methanogenic and denitrifying bacterial consortia, indole was degraded, proceeding through a two-step hydroxylation pathway, yielding oxindole and isatin subsequent to cleavage between the C-2 and C-3 atoms on the pyrrole ring of indole [Berry et al., 1987; Gu and Berry, 1991; Gu and Berry, 1992; Madsen et al., 1988; Madsen and Bollag, 1989]. D. indolicum, the only sulfate-reducing bacterium

Kinetics and Metabolic Pathway of Melatonin Biodegradation…

301

isolated so far, could use indole as the sole carbon source and electron donor [Bak and Widdel, 1986]. When authentic oxindole and isatin standards were used for identifying the intermediates which appeared in the culture, the unidentified peak did not match those of either oxindole or isatin. Furthermore, Ps. aeruginosa Gs could not utilize oxindole or isatin as the sole carbon and energy source, suggesting that a new biochemical pathway for indole degradation by Ps. aeruginosa Gs exists.

4. CONCLUSION A microorganism capable of using melatonin as the sole sources of carbon and energy under aerobic conditions was isolated. It was identified as Shewanella putrefaciens. The experimental results showed that 50 mg l−1 melatonin could be degraded within 2 d under the optimal condition (pH 7.0, salinity 15‰ and temperature at 37 °C). The process of MLT biodegradation was monitored by high-performance liquid chromatography with ultra-violet detection. Kinetics study revealed that the biodegradation of melatonin by Shewanella putrefaciens followed the first-order reaction kinetics when initial melatonin concentration was in the range of 25–150 mg l−1. The major metabolites of melatonin biodegradation were identified by high-performance liquid chromatography and gas chromatography-mass spectrometry, and a preliminary metabolic pathway of melatonin was proposed. The results are helpful to understand behavior of MLT in environment, and also could be useful for the bioremediation of MLT-contaminated site, such as the wetland of the Mai Po Natural Reserve in Hong Kong.

5. REFERENCES Bak, F. and Widdel, F. (1986). Anaerobic degradation of indolic compounds by sulfatereducing enrichment cultures, and description of Desulfobacterium indolicum gen. nov. sp. nov. Arch. Microbiol., 146:170-176. Berry, D.F., Francis, A.J. and Bollag, J.M. (1987). Microbial metabolism of homocyclic and heterocyclic aromatic compounds under anaerobic conditions. Microbiol. Reviews, 51: 43-59. Chang, B.V., Chiang, F. and Yuan, S.Y. (2005). Anaerobic degradation of nonylphenol in sludge. Chemosphere, 59: 1415-1420. Claus, G. and Kutzner, H.J. (1983). Degradation of indole by Alcaligenes spec. Syst. Appl. Microbiol., 4: 169-180. Doukyu, N. and Aono, R, (1997). Biodegradation of indole at high concentration by persolvent fermentation with Pseudomonas sp. ST-200, Extremophiles, 1: 100–105. Gibson, D.T. (1984). Microbial Degradation of Organic Compounds. New York: Marcel Dekker. Gu, J.D. and Berry, D.F. (1991). Degradation of substituted indoles by an indole-degrading methanogenic consortium. Appl. Environ. Microbiol., 57: 2622-2627. Gu, J.-D. and Berry, D.F. (1992). Metabolism of 3-methylindole by a methanogenic consortium. Appl. Environ. Microbiol., 58: 2667-2669.

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Gu, J.D., Fan, Y.Z. and Shi, H.C. (2002). Relationship between structures of substituted indolic compounds and their degradation by marine anaerobic microorganisms. Mar. Pollut. Bull., 45: 379-384. Jobling, S., Nolan M., Tyler, C.R., Brighty, G. and Sumpter, J.P. (1998). Widespread sexual disruption in wild fish. Environ. Sci. Technol., 32: 2498-2506. Kaiser, J.P., Feng, Y. and Bollag, J.M. (1996). Microbial metabolism of pyridine, quinoline, acridine and their derivatives under aerobic and anaerobic conditions. Microbiol. Reviews, 60: 483-498. Kolar, J. and Machackova, I. (2005). Melatonin in higher plants: occurrence and possible functions. J. Pineal Res., 39: 333-341. Luboshitzky, R., Shen-Orr, Z., Nave, R., Lavi, S. and Lavie, P. (2002). Melatonin administration alters semen quality in healthy men. J. Androl., 23: 572-578. Madsen, E.L. and Bollag, J.M. (1989). Pathway of indole metabolism by denitrifying microbial community. Arch. Microbiol., 151: 71-76. Reiter R.J. (1991). Pineal melatonin – cell biology of its synthesis and its physiological interactions. Endocr. Rev., 12: 151-180. Madsen, E.L., Francis, A.J. and Bollag, J.M. (1988). Environmental factors affecting indole metabolism under anaerobic conditions. Appl. Environ. Microbiol., 54: 74-78. Oshima, T. Kawai, S. and Egami, F. (1965). Oxidation of indole to indigotin by Pseudomonas indoloxidans. J. Biochem., 58: 259-263. Tan, D.X., Manchester, L.C., Hardeland, R., Lopez-Burillo S., Mayo, J.C., Sainz, R.M. and Reiter, R.J. (2003). Melatonin: a hormone, a tissue factor, an autocoid, a paracoid, and an antioxidant vitamin. J. Pineal Res., 34: 75-78. Tanyildizi, S., Bozkurt, T., Ciftci, O. and Sakin, F. (2006). In vitro effects of melatonin on hyaluronidase activity and sperm motility in bull semen. Turk. J. Vet. Anim. Sci., 30: 8993. Wang, Y.Y., Fan, Y.Z. and Gu, J.D. (2004). Dimethyl phthalate ester degradation by two planktonic and immobilized bacterial consortia. Int. Biodeter. Biodegr., 53: 93-101. Xu, X.R., Gu, J.D., Li, H.B. and Li, X.Y. (2005). Kinetics of di-n-butyl phthalate degradation by a bacterium isolated from mangrove sediment. J. Microbiol. Biotechnol., 15: 946-951. Xu, X.R., Li, H.B. and Gu, J.D. (2006). Elucidation of n-butyl benzyl phthalate biodegradation using high-performance liquid chromatography and gas chromatographymass spectrometry. Anal. Bioanal. Chem., 386: 370-375. Xu, X.R., Li, H.B. and Gu, J.D. (2007). Metabolism and biochemical pathway of n-butyl benzyl phthalate by Pseudomonas fluorescens B-1 isolated from a mangrove sediment. Ecotox. Environ. Safe., (In press and available online 12 February 2007). Yamada, K., Maruyama, K., Mogami, S., Miyagawa, N. and Tsuboi, M. (1992). Influence of melatonin on reproductive behavior in male rats. Chem. Pharm. Bull., 40: 2222-2223. Yilmaz, B., Kutlu, S., Mogulkoc, R., Canpolat, S., Sandal, S., Tarakci, B. and Kelestimur, H. (2000). Melatonin inhibits testosterone secretion by acting at hypothalamo-pituitarygonadal axis in the rat. Neuroendocrinol. Lett., 21: 301-306. Yin, B., Huang, L.M. and Gu, J.D. (2006). Biodegradation of 1-methylindole and 3methylindole by mangrove sediment enrichment cultures and a pure culture of an isolated Pseudomonas aeruginosa Gs. Water Air Soil Pollut., 176: 185-199. Young, L.L. and Cerniglia, C.E. (1995). Microbial Transformation and Degradation of Toxic Organic Chemicals. New York: John Wiley.

INDEX

A AC, 35, 75, 186 accelerator, 162 access, 247 acclimatization, 138, 168, 276 acetic acid, 19 acetone, viii, 49, 181 acetonitrile, 179, 277 acid, 4, 7, 16, 17, 18, 19, 20, 21, 24, 25, 26, 30, 34, 35, 37, 40, 41, 55, 60, 65, 66, 72, 81, 85, 88, 166, 185, 191, 192, 193, 197, 200, 206, 208, 213, 219, 224, 225, 230, 231, 232, 233, 234, 235, 237, 245, 246, 248, 256, 260, 261, 270, 300 activated carbon, ix, 78, 95, 97, 98, 170, 180, 275 activation, 30, 33, 35, 38, 40, 76, 99, 180, 253 active oxygen, 20 active site, 41 adaptability, 97 adaptation, 132, 138 additives, 140, 143, 168, 241, 247, 255, 286 adenine, 130 adenocarcinoma, 40 adenosine, 250 adhesion, 8, 256 adhesives, 50 adipose, 65, 72 adipose tissue, 65, 72 administration, 62, 72, 74, 302 ADP, 250 adsorption, viii, ix, xi, 7, 50, 77, 78, 79, 84, 93, 94, 95, 96, 97, 98, 100, 145, 163, 170, 260, 268, 275 adults, 68, 73 aerospace, 50 agar, 219, 276, 291, 293 agent, 22, 25, 52, 82, 164, 235, 242 agglutination, 45

aggregates, 132, 145 aggregation, 13 aging, 128, 149, 156, 241 agonist, 53 agriculture, viii, ix, xi, 77, 78, 95, 218, 273 alanine, 230 alcohol, 67, 81 alcohols, 61 alfalfa, 10, 28, 39 algae, 26, 33, 60, 64, 72 alginate, viii, 2, 5, 25, 26, 27, 33, 34, 35, 36, 37, 40, 42, 43, 44, 47, 260 algorithm, 124 alkaloids, 274 alternative, 7, 89, 91, 92, 102, 105, 110, 191, 209, 212, 218, 246, 247, 249, 253, 275 alternatives, x, 218, 239 alters, 302 ambiguity, 177 ambivalence, 240 amendments, 233 amide, 246 amines, 11, 29, 61 amino acid, 19, 37, 221, 222, 225, 230, 231, 232, 233, 234, 246 amino acids, 221, 222, 225, 230, 231, 232, 233, 234, 246 ammonia, 3, 37, 44, 91, 167, 168, 225, 232, 234 ammonium, vii, 46, 82, 85 amorphous polymers, 209 amphibians, viii, 49, 55, 65 amyotrophic lateral sclerosis, 161, 184 anaerobe, 17, 20, 35, 47 anaerobic bacteria, 57, 126 androgen, 62, 74 animals, xii, 67, 81, 289, 290 anion, 22, 46, 222 annealing, 124

304

Index

antagonism, 219, 223 antibiotic, 46 antibiotic resistance, 46 anticoagulant, 47 antimicrobial protein, 2, 27 antioxidant, 302 antitumor, 2, 11 antiviral, 10, 166 AP, 33, 36 apatites, 271 aqueous solutions, 230 aquifers, 103, 105, 106, 108, 110, 111, 112, 113, 115, 117, 119, 120, 122, 123, 125 argon, 192 Aristotle, 189 aromatic compounds, 104, 111, 274, 281, 301 aromatic hydrocarbons, ix, 105, 106, 119, 123, 125, 127, 128, 130, 281, 285 aromatics, ix, 101, 104, 106, 111, 284, 300 Aspergillus terreus, 56, 58 assessment, 75, 100, 116, 117, 118, 119, 124, 125, 237, 268 assumptions, 108, 112, 113, 115, 138 atoms, 105, 300 ATP, 249, 250, 256 attachment, 7, 8 attacks, 161 attention, x, xi, 16, 25, 106, 140, 159, 189, 191, 241, 273, 274 availability, xi, 89, 104, 134, 225, 230, 236, 239, 260 azo dye, 161

B BAC, 95 Bacillus, 12, 17, 18, 23, 24, 26, 27, 29, 32, 33, 36, 37, 40, 41, 42, 45, 222, 286 Bacillus subtilis, 29, 45, 222 bacteria, vii, viii, xii, 7, 8, 11, 17, 18, 20, 25, 29, 30, 31, 42, 49, 55, 57, 58, 65, 69, 98, 104, 117, 128, 130, 131, 133, 134, 135, 138, 139, 166, 184, 219, 220, 228, 235, 247, 255, 275, 283, 285, 289, 290, 291, 295, 300 bacterial cells, 7, 8, 29, 55, 244 bacterial strains, 22, 85, 250 bacteriophage, 26 bacteriostatic, 218 bacterium, xii, 5, 16, 18, 21, 22, 23, 24, 26, 32, 35, 37, 39, 41, 42, 43, 44, 46, 69, 71, 72, 99, 166, 192, 220, 256, 289, 293, 300, 302 banks, 171 barley, 16, 41, 44 beef, 219

behavior, xii, 76, 124, 138, 145, 196, 197, 198, 203, 204, 210, 213, 218, 281, 290, 301, 302 benefits, 78, 114 benign, ix, 127 benzene, vii, x, 61, 104, 106, 110, 120, 121, 122, 123, 125, 128, 130, 159, 160, 172, 173, 174, 175, 176, 177, 178, 179, 182, 274, 275, 287 bile, 62, 63, 69 binary blends, 209 binding, 7, 8, 10, 29, 33, 34, 46, 51, 66, 145, 236, 243, 270 bioaccumulation, viii, 49, 50, 64, 66, 78 bioactive materials, 260, 269 bioassay, 224, 277 bioavailability, ix, 72, 78, 93, 95, 104, 127, 128, 132, 133, 141, 144, 145, 149, 212 biochemistry, 34, 240 biocide activity, 166 biodegradability, x, xi, xii, 55, 57, 58, 75, 104, 170, 191, 199, 208, 239, 240, 254, 255, 256, 259, 276, 285, 286, 289, 290 biodegradable, vii, x, xi, 152, 189, 190, 191, 208, 209, 213, 239, 240, 253, 254, 255, 259, 260, 268 biodegradable materials, 208, 240, 253 biofilm formation, 256 biokinetics, 120 biological activity, 28, 30, 31, 32, 133 biological control, 2, 25 biological control agents, 2 biological macromolecules, 2 biological responses, 270 biomarkers, 119, 126 biomass, vii, xi, 79, 92, 94, 95, 109, 111, 135, 137, 139, 146, 152, 218, 220, 224, 235, 237, 239, 241, 248, 249, 250, 255, 256, 277, 279 biomass growth, 109 biomaterials, 263, 268, 269, 271 biomolecule, 43, 268 biomolecules, 246 bioremediation, ix, xi, xii, 88, 91, 97, 118, 119, 121, 122, 123, 124, 127, 128, 129, 130, 132, 135, 136, 140, 141, 142, 143, 144, 148, 149, 152, 153, 157, 274, 276, 280, 282, 290, 301 biosynthesis, 7, 28, 30, 33, 43, 47 biotechnology, 35 biotic, 78, 240, 245 birds, 68 bisphenol, 50, 53, 56, 63, 64, 66, 67, 68, 69, 70, 71, 72, 73, 74, 75, 76 black tea, 186 bleaching, 52 blends, x, 189, 209, 210, 211, 213, 240 blocks, 25, 203

Index blood, 63, 100 blood flow, 63 body fluid, 261, 270 bonds, 12, 59, 61, 110, 198, 245, 246 bone grafts, 260 bone marrow, 264, 265, 266, 267, 268, 269, 270, 271 bone remodeling, 263, 268 bone resorption, 268 bone tumors, 269 branching, 46 breakdown, 18, 41, 65, 97, 105, 170 breast cancer, 71, 73 breeding, 67 brominated flame retardants, 68 buffer, 46, 198, 199, 204 building blocks, 9, 128, 130 burn, 130 butyl ether, 105, 106, 119 by-products, 52, 68, 87, 121, 191, 218

C Ca2+, 58 cabbage, 37 cadmium, 71, 122 calcium, xi, 17, 47, 143, 259, 260, 261, 263, 267, 268, 269, 270, 271 calvaria, 261, 270 canals, 84 cancer, 122, 166 candidates, 75 capillary, 185, 237, 292 carbohydrate, 7, 8, 10, 15, 22, 31, 219 carbohydrates, viii, 1, 8, 16, 38 carbon, vii, ix, xi, xii, 18, 55, 58, 60, 78, 79, 84, 85, 87, 89, 90, 91, 92, 93, 95, 104, 109, 110, 115, 119, 120, 123, 124, 125, 139, 142, 143, 167, 168, 170, 180, 182, 192, 208, 218, 222, 235, 236, 239, 242, 250, 252, 253, 255, 274, 276, 278, 279, 280, 281, 283, 285, 289, 290, 291, 293, 300, 301 carbon atoms, 104 carbon dioxide, vii, xi, 218, 222, 239, 250, 255 carbonyl groups, 245 carboxylic acids, 61 carcinogen, 128 carcinogenicity, 81, 104, 185 carcinogens, 104 carcinoma, 46 carrageenan, viii, 2, 20, 21, 35, 39, 41, 47 carrier, 222, 292 case study, 98 cast, 209 casting, 210

305

catalyst, 160, 192, 193 catalysts, 191, 208 catalytic hydrogenation, 190 catalytic properties, 19 catfish, 238 cation, 222 cattle, 236 C-C, 34 cDNA, 47 cell, vii, xi, xii, 1, 2, 3, 4, 5, 6, 7, 8, 9, 10, 11, 15, 16, 18, 20, 22, 25, 27, 29, 30, 31, 33, 34, 36, 38, 40, 42, 44, 45, 46, 57, 60, 71, 73, 84, 89, 90, 91, 98, 184, 186, 194, 219, 224, 229, 235, 249, 260, 263, 264, 267, 268, 271, 274, 302 cell adhesion, 40 cell culture, 186, 264, 268 cell death, 8, 36 cell differentiation, xi, 260, 268 cell growth, xii, 11, 22, 25, 29, 57, 84, 89, 90, 91, 274 cell line, 184, 267 cell lines, 184 cell membranes, 5, 34 cell metabolism, 98 cell surface, 7, 8, 11, 27, 29, 30 cellulose, 11, 15, 16, 42, 46, 143, 219, 253 ceramic, xi, 220, 259, 266, 267, 269 ceramics, 260, 269, 271 cereals, 33 certainty, ix, 102, 103, 114 certification, 255 chain molecules, 247 chain scission, 208 charge density, 7 chemical bonds, 245 chemical composition, 261 chemical degradation, viii, 2, 9, 58, 96 chemical properties, 139 chemical structures, 20, 209 chemometrics, 255 chicken, 62 chimpanzee, 63, 75 China, 41, 51, 52, 73, 76 chitin, viii, 2, 3, 5, 11, 13, 14, 29, 30, 38, 39, 42, 45, 46, 47 chitosan, viii, 2, 3, 5, 7, 11, 12, 13, 14, 22, 29, 31, 33, 34, 35, 37, 38, 39, 40, 41, 42, 43, 44, 45, 46, 47, 143, 158, 260 chloride, vii, xii, 72, 274, 275, 276, 278, 279, 281, 284 chlorinated aliphatics, 123 chlorination, 68, 81 chlorine, 52, 105

306

Index

chloroform, 125, 185, 206, 209, 210, 212, 247 chromatographic technique, 248 chromatography, xii, 21, 46, 69, 70, 72, 82, 97, 100, 123, 160, 177, 185, 208, 220, 247, 248, 252, 281, 290, 292, 298, 301, 302 chromatography analysis, 281 circulation, 82, 218, 234, 252 classes, 145, 149 classification, 37, 99 cleaning, 93, 140 cleavage, 5, 12, 81, 106, 128, 183, 284, 300 climate change, 114 cloning, 24, 33, 37, 47 cluster analysis, 19 CO2, vii, 55, 57, 60, 109, 112, 137, 170, 192, 223, 224, 227, 229, 240, 250, 251, 252, 253 coagulation, 182 coal, xi, 102, 103, 120, 170, 273 coal tar, 102, 103, 120 coatings, 50, 271 collagen, 260, 262 colloids, 131 colonization, xi, 16, 239, 244 combined effect, 116 community, 78, 89 compatibility, 209 competition, 223 complement, 18, 93 complementarity, 183 complexity, 44, 108, 120, 145 complications, x, 239 components, vii, 3, 16, 20, 30, 35, 38, 179, 209, 210, 263, 264, 276 composition, 9, 13, 39, 87, 107, 200, 208, 209, 219, 222, 225, 230, 231, 233, 234, 250 compost, 76, 208, 219, 225, 236, 237, 238, 250, 253 composting, 219, 236, 237, 251, 254, 255 compounds, vii, ix, x, xi, xii, 9, 39, 58, 60, 61, 66, 67, 70, 72, 75, 78, 80, 84, 87, 88, 91, 96, 101, 102, 103, 104, 105, 106, 110, 115, 117, 121, 124, 126, 127, 128, 130, 138, 140, 141, 142, 143, 150, 159, 160, 162, 163, 164, 166, 167, 168, 170, 178, 180, 183, 184, 185, 218, 225, 228, 239, 246, 247, 248, 253, 255, 259, 260, 261, 267, 273, 274, 275, 276, 277, 278, 280, 281, 282, 283, 284, 285, 286, 287, 292, 298, 301, 302 concentration, viii, x, xii, 13, 57, 64, 74, 77, 78, 82, 84, 85, 91, 92, 93, 94, 95, 96, 105, 109, 111, 112, 113, 115, 118, 127, 136, 139, 145, 146, 147, 148, 149, 152, 163, 164, 166, 167, 168, 182, 217, 219, 221, 222, 224, 225, 229, 233, 234, 237, 253, 275, 277, 278, 279, 282, 283, 284, 289, 290, 292, 293, 294, 297, 301

conceptual model, 112 conductivity, 83, 93, 109 configuration, 252 confinement, 96, 97 conjugation, 62 connective tissue, 263, 264 consensus, 268 constant rate, 295 constituent groups, 18 constraints, 144 consulting, 254 consumer protection, 100 consumption, 87, 94, 200, 211, 223, 283, 284 contact dermatitis, 183 contact time, 128, 144 contaminant, x, 84, 95, 98, 105, 106, 107, 109, 110, 111, 112, 115, 116, 117, 118, 121, 122, 126, 127, 129, 132, 137, 142, 144, 145, 146, 147, 148, 149, 150, 152, 274 contaminants, vii, 73, 102, 103, 106, 107, 116, 117, 119, 120, 123, 124, 125, 137, 140, 144, 145, 146, 148, 149, 284, 285 contaminated soils, 102, 123, 125 contamination, viii, 49, 50, 51, 53, 64, 65, 70, 75, 77, 78, 81, 82, 84, 95, 96, 102, 106, 107, 108, 112, 114, 116, 118, 122, 139, 140, 163, 274, 284, 290 continuity, 263 control, 2, 7, 13, 14, 36, 37, 70, 79, 84, 94, 96, 140, 143, 218, 221, 227, 229, 232, 236, 237, 238, 241, 242, 251, 252, 253, 268, 277, 278, 296 conversion, vii, xi, 11, 70, 104, 137, 193, 224, 237, 259, 268, 269, 270 cooling, 196, 212 copolymers, x, 189, 190, 199, 200, 201, 202, 203, 204, 205, 207, 210, 213, 260 copper, 72 coral reefs, 70 corn, 161, 170 correlation, 110, 182, 183, 296 correlation coefficient, 296 correlations, 178 cosmetics, 25 cost effectiveness, 141 costs, 102, 140 cotton, 236 cotyledon, 22, 26, 28, 30, 31 coumarins, 274 coupling, 67, 69, 120, 144, 146 coverage, 243, 244 covering, 242 criticism, 253 crop production, 233 crops, 78, 161

Index crude oil, 112, 117, 121, 125 crystal growth, 261 crystal structure, 190, 194, 202 crystalline, 196, 197, 200, 204, 212, 260, 271 crystallinity, 199, 200, 203, 204, 211, 247, 260 crystallites, 200 crystallization, 41, 196, 197, 202, 209 crystals, xi, 194, 200, 204, 205, 209, 244, 259, 260, 261, 269, 270 cultivation, 220, 228, 230, 250, 281, 282 culture, ix, 9, 17, 19, 24, 37, 66, 77, 78, 79, 85, 87, 88, 90, 91, 92, 93, 94, 95, 96, 97, 98, 99, 100, 167, 172, 185, 219, 223, 229, 237, 267, 268, 277, 279, 280, 281, 282, 283, 284, 290, 291, 292, 293, 297, 298, 300, 302 culture conditions, 185, 284 current limit, 141 cyanocobalamin, 122 cycles, 253 cystic fibrosis, 35, 42 cysts, 256, 278 cytochrome, 57, 62, 63, 68, 72, 73, 99 cytosine, 130

D dairy, 226, 237 database, 100, 235, 298 DD, 75 death, 7, 29, 166 death rate, 29 decay, 109 decomposition, vii, 69, 192, 197, 198, 203, 204, 205, 225, 236, 252 decomposition reactions, 192 decomposition temperature, 198 decontamination, viii, ix, 77, 78, 84, 96, 98 defects, xi, 200, 242, 259, 261, 262, 263, 270 defense, vii, 1, 2, 3, 5, 6, 8, 9, 18, 20, 22, 27, 30, 32, 33, 34, 35, 38, 39, 40, 41, 43, 44 defense mechanisms, vii, 1, 2, 5, 18, 33, 43 deficiency, 232 definition, 102, 103 deformation, 246 degradation pathway, 80, 179, 281, 284, 300 degradation process, 78, 97, 241 degradation rate, 86, 89, 90, 91, 109, 110, 111, 125, 139, 142, 155, 168, 198, 199, 201, 213, 283, 295, 296, 297 degree of crystallinity, x, 189, 190, 196, 199, 201, 240 dehalogenation, 53, 121, 125 demand, 212, 220, 222, 229, 237, 255

307

denitrification, 135, 163 denitrifying, 237, 300, 302 density, 13, 81, 108, 139, 144, 146, 149, 152, 255 dentin, 261 Department of Energy, 153 depolymerization, viii, 2, 9, 11, 23, 37, 39, 42 deposition, 5 depression, 200, 204, 209, 210, 211 derivatives, x, 25, 37, 52, 53, 54, 67, 74, 89, 159, 160, 161, 162, 163, 165, 167, 168, 170, 182, 185, 186, 302 dermatitis, 166 desiccation, 250 desorption, 95, 97, 108, 132, 140, 145, 148 destruction, vii, 103, 143, 275, 276 detection, xii, 40, 83, 84, 86, 91, 94, 219, 248, 249, 252, 256, 289, 290, 292, 298, 301 detoxication, viii, 49, 66 DG, 45, 270 diesel fuel, 126 differential equations, 138 differentiation, 198, 267, 268 diffusion, 106, 107, 108, 110, 111, 113, 126, 128, 131, 144, 145, 148, 252 diffusion process, 145 diffusivity, 146, 152 digestion, 38, 46 dimer, 29 dimethylsulfoxide, 247 dioxin, 66 discharges, 274 discounting, 114 discrimination, 255 discs, 271 dispersion, 82, 106, 109, 110, 119, 125 disposition, 73 dissociation, 81 dissolved oxygen, 221 distillation, 110, 192 distribution, vii, 1, 3, 7, 14, 27, 29, 30, 31, 34, 60, 68, 82, 118, 121, 139, 248 diversity, viii, 2, 46, 77, 97 DNA, 7, 32, 64, 66, 128, 130, 220 donors, 103, 105 dosage, 212 draft, 158 drainage, 51, 80, 82, 84 drinking water, 70, 81, 102, 115, 187, 274 drug delivery, x, 189, 212 drug delivery systems, x, 189, 212 drug interaction, 70 drugs, x, 190, 212 DSC, 196, 199, 203, 204, 209, 210

308

Index

DSM, 99 dumping, 217 duration, 141 dyes, 242, 243, 249

E ecology, 117, 150, 286 economics, 114 ecosystem, 140, 290 effluent, viii, ix, 49, 50, 51, 65, 70, 72, 78, 92, 95, 97, 98, 168, 170, 217, 226, 237, 285 effluents, viii, ix, xi, 51, 77, 78, 93, 95, 96, 97, 100, 163, 164, 165, 170, 185, 218, 273, 275, 282, 285, 286 egg, 42 eggs, 12, 68 electrical conductivity, 95 electrodes, 287 electron, xi, 57, 58, 89, 91, 103, 105, 110, 111, 112, 120, 125, 128, 130, 132, 134, 135, 136, 139, 142, 144, 146, 148, 152, 239, 242, 244, 292, 301 electron microscopy, xi, 239, 242, 244 electrophoresis, 185, 222, 237 elongation, 198, 203, 224, 230 emission, 67, 184, 233 encoding, 2, 24, 33, 43 endocrine, viii, 49, 50, 66, 69, 70, 72, 74, 75, 125 endocrine-disrupting chemicals, 66, 74, 75 energy, vii, xii, 2, 18, 55, 58, 72, 85, 87, 89, 92, 95, 128, 130, 136, 170, 237, 256, 275, 279, 280, 289, 290, 291, 292, 293, 300, 301 entrapment, 108, 109 environment, viii, x, xii, 2, 49, 50, 51, 53, 55, 58, 60, 65, 67, 69, 75, 77, 78, 81, 82, 84, 99, 125, 128, 130, 132, 136, 140, 141, 142, 143, 159, 160, 162, 163, 164, 166, 178, 179, 182, 185, 190, 208, 209, 248, 254, 274, 275, 282, 284, 286, 289, 290, 300, 301 environmental characteristics, 130 environmental conditions, 97, 99, 128, 132, 170, 190 environmental contamination, viii, 77, 78, 84, 100 environmental factors, 128, 150 Environmental Protection Agency, 111, 128, 274, 287 environmental standards, 114 enzymatic activity, 13 enzyme, 6, 11, 13, 14, 16, 17, 19, 24, 26, 27, 34, 38, 41, 45, 47, 61, 68, 72, 75, 131, 133, 174, 199, 249, 275, 284, 287, 297 enzymes, viii, 1, 3, 5, 6, 9, 11, 12, 13, 14, 16, 17, 18, 21, 24, 26, 27, 31, 33, 36, 37, 41, 42, 44, 45, 46,

58, 59, 60, 61, 62, 63, 66, 72, 73, 74, 132, 174, 200, 250, 256, 284, 286, 297, 300 EPA, 130, 158 epoxy, viii, 49, 50 epoxy resins, viii, 49, 50 equilibrium, 95, 107, 145, 147, 149, 196, 209 equipment, 218 equity, 114 erosion, 150, 200, 202, 211 erythrocytes, 45 Escherichia coli, 24, 29, 45, 47, 237, 254 ESI, 35, 160, 174 ester, 20, 34, 88, 200, 205, 206, 211, 213, 246, 302 ester bonds, 200, 211 esters, 85, 198, 208, 245 estimating, 119, 150 estradiol, 73 estrogen, 53, 73, 75 ethanol, 45, 81, 118, 179, 242 ethers, 118, 198 ethyl acetate, 81, 90 ethylene, x, 5, 189, 190, 196, 207, 209, 211, 213 ethylene glycol, 207, 213 ethylene oxide, 211 evaporation, 210 evidence, 104, 108, 111, 115, 117, 121, 124, 185, 209, 248, 249, 260, 261, 269, 278 evolution, vii, 40, 41, 99, 168, 197 excision, 269 exclusion, 208, 247, 248 excretion, 61, 63, 69, 71 exopolysaccharides, 3, 31 exothermic, 218 experimental condition, 250, 263 expertise, xi, 239, 240 exploitation, 97 explosives, 104, 124 exposure, viii, 6, 49, 50, 62, 64, 68, 70, 97, 105, 113, 114, 115, 122, 128, 130, 133, 166, 182 Exposure, 155 extraction, 18, 20, 82, 97, 100, 176, 220, 247, 249, 256, 274, 275, 292, 296, 298 extrapolation, 67

F FAD, 168 failure, 140 fairness, 114 false positive, 242 family, x, 7, 16, 17, 19, 20, 24, 29, 32, 34, 37, 39, 40, 43, 44, 74, 85, 86, 159, 160, 161, 171, 249 family members, 24

Index farming techniques, 121 farms, 236 fatty acids, 182, 246 feces, 62, 63 feet, 142 femur, 263, 264, 266, 267 fermentation, 191, 218, 229, 237, 301 ferric ion, 72 ferrous ion, 34 fertility, 225, 230 fertilization, 237 fertilizers, 143 fibers, 190, 208 fibrin, 260 fibrous tissue, 260 film, 132, 137, 210, 241, 242, 243, 244, 245, 248, 250, 275, 286, 292 film thickness, 292 films, 32, 198, 199 filtration, 170 fish, viii, 46, 49, 55, 61, 62, 64, 65, 66, 70, 100, 128, 130, 217, 220, 222, 234, 235, 237, 302 fisheries, 217, 237 fixation, 244 fixed bed reactors, 275, 286 flame, 53, 66, 67, 82 flame retardants, 67 flavonoids, 274 flocculation, 182 flood, 73, 79 flow field, 124 fluid, 62, 70, 146, 152 fluid extract, 70 fluidized bed, 142 fluorescence, xi, 239, 243 foams, 220 folic acid, 58 food, viii, ix, 2, 25, 49, 50, 64, 65, 127, 128, 218 formaldehyde, 242 formamide, 21 fossil, 274, 284 fouling, 13, 14 Fourier, 245, 255 fragmentation, 2, 22, 28, 292 free energy, 136 freshwater, 50, 53, 61, 64, 65, 68, 121, 274, 297 Freundlich isotherm, 145 fruits, 35, 41, 60, 70, 74 FTIR, 241, 245, 246, 256 fuel, 105, 120, 124, 274, 284 fumigation, 183 fungal infection, 3

309

fungi, viii, xii, 3, 11, 12, 16, 19, 25, 28, 29, 32, 39, 49, 55, 58, 65, 67, 71, 73, 85, 131, 160, 161, 255, 275, 287, 289, 290 fungus, xi, 9, 33, 38, 47, 66, 72, 126, 176, 177, 185, 257, 274 fusion, 196, 210

G gas phase, 230 gases, 110, 184, 222 gasoline, 118 gender, 74 gender differences, 74 gene, 5, 24, 31, 32, 33, 34, 35, 36, 37, 38, 40, 41, 43, 45, 73, 85, 86, 87, 222, 271 gene expression, 5, 31, 36, 38, 73, 271 generation, 2, 235 genes, 3, 6, 30, 35, 40, 43, 44, 220 germination, 37, 219, 221, 224, 236, 237 Gibbs free energy, 135 glass, 192, 196, 197, 201, 203, 206, 208, 209 glass transition, 196, 197, 201, 203, 206, 208, 209 glass transition temperature, 196, 197, 201, 203, 206, 208, 209 glucan, vii, 1, 3, 5, 6, 9, 10, 16, 22, 24, 28, 30, 33, 34, 38, 39, 41, 43, 44, 46, 47 glucose, 3, 4, 10, 18, 20, 24, 28, 33, 58, 191, 219, 276 glucoside, 10, 28, 29, 33, 34, 60 glutamic acid, 230 glutathione, 64 glycerol, 191 glycine, 33, 47, 230, 237 glycol, x, 189, 190, 193 glycoside, 17, 32, 40, 41 glycosylation, 60 goals, 142 gold, 260 government, 102, 240 GPC, 198, 203, 209 gracilis, 56 gradient formation, 235 granules, 266 grass, 79 grazing, 131 ground water, viii, 77, 84, 98, 162 groundwater, 102, 103, 105, 106, 107, 110, 112, 113, 114, 115, 116, 117, 118, 120, 122, 123, 124, 125, 126, 143, 285, 286 groups, vii, 1, 3, 7, 11, 16, 18, 19, 20, 22, 29, 31, 45, 50, 178, 193, 196, 198, 203, 245, 246, 255, 262, 265, 268

310

Index

growth, vii, xii, 2, 18, 22, 23, 25, 29, 31, 33, 35, 37, 38, 46, 58, 70, 85, 87, 89, 90, 91, 92, 94, 118, 138, 146, 152, 167, 168, 184, 191, 202, 218, 219, 221, 222, 224, 225, 228, 232, 235, 236, 245, 248, 250, 254, 261, 267, 268, 276, 279, 281, 289, 293, 296, 297, 300 growth factor, 236 growth factors, 236 growth rate, 146, 152, 228, 293 guanine, 130 guard cell, 40 guidelines, 123

H habitat, 64 half-life, 61, 65, 83, 171, 296 halogens, 53 harm, 102 harmful effects, 2, 81 hazardous substances, 111 hazards, 128 health, ix, 2, 78, 127, 130, 150, 160, 187, 274 heat, 31, 137, 168, 196, 210, 246 heat conductivity, 246 heating, 192, 196, 198, 203, 205, 209 heating rate, 196, 198 heavy metals, 103, 236 height, 146, 152 helium, 222, 292 hematoxylin-eosin, 262 heme, 58 hemicellulose, 15, 19 hepatocytes, 71 hepatotoxicity, 64 herbicide, 79, 81, 82, 84, 85, 87, 89, 91, 92, 93, 95, 96, 97, 98, 99, 100, 161, 170, 171, 176, 182, 185 heterogeneity, 35, 108, 109, 111, 112, 113 hexane, 81 high-performance liquid chromatography, xii, 289, 290, 292, 298, 301, 302 HIV, 47 homogeneity, 10 homopolymers, 200, 213 Honda, 68, 270 hormone, xii, 5, 289, 290, 302 host, 5, 7, 8, 20, 30, 31, 32, 33, 38, 39, 260 HPLC, xii, 22, 86, 87, 90, 99, 137, 172, 173, 274, 277, 280, 291, 292, 298 human exposure, 123 human wastes, viii, 49, 50, 51, 65 humic substances, 76, 110, 111, 120 humidity, 82, 221

humus, vii, 219 hybrid, 53, 124 hydrocarbons, ix, 101, 104, 105, 107, 112, 116, 119, 121, 124, 125, 182, 191 hydroformylation, 191 hydrogen, 15, 34, 65, 119, 121, 275, 287, 290, 292, 296, 297 hydrogen bonds, 15 hydrogen peroxide, 34, 65, 275, 287, 290, 292, 296 hydrology, 112 hydrolysates, viii, 2, 14, 22, 25, 31, 192 hydrolysis, 9, 11, 12, 13, 14, 16, 18, 20, 21, 24, 28, 31, 33, 34, 37, 39, 42, 43, 44, 46, 47, 106, 198, 199, 200, 201, 202, 204, 205, 208, 209, 210, 211, 213, 219, 268 hydrophilicity, 7, 30 hydrophobicity, 30, 204 hydroquinone, 276, 280, 286 hydroxide, 250 hydroxides, 182 hydroxyapatite, xi, 259, 260, 261, 264, 265, 266, 267, 269, 270 hydroxyl, 20, 60, 65, 203, 274, 281 hydroxyl groups, 20, 274 hypothesis, 8, 103, 105, 197, 202 hypoxia, 253 hysteresis, 108

I ICM, 35 identification, 37, 66, 71, 85, 113, 174, 175, 176, 177, 185, 222, 236, 237, 291, 292, 298 identity, 87 images, 243, 244 immobilization, 13, 14, 285 immobilized enzymes, 13 immunity, 32 implants, 263, 264, 265, 266, 271 implementation, viii, 77, 78, 88, 93, 96, 97 in situ, x, 102, 116, 119, 120, 121, 124, 126, 127, 128, 134, 140, 141, 142, 143, 144, 152, 160, 172, 173, 174, 247 in vitro, xi, 22, 40, 42, 43, 64, 260, 261, 267, 268 in vivo, xi, 64, 74, 260, 261, 268, 271 inclusion, 261, 270 incubation period, 91, 93, 137, 248 incubation time, 137, 170, 171, 172, 277 independent variable, 147 indication, 203, 210, 225, 250 indicators, xi, 118, 239 indices, 88, 260 indigenous, 128, 140

Index inducer, 104, 284 induction, 3, 5, 7, 28, 30, 37, 40, 41, 44, 47, 64, 197, 267 induction period, 197 industrial application, x, 13, 159, 160 industrial production, 218 industrial wastes, 51, 118, 274 industry, 85, 102, 161, 208, 218, 226, 235, 286 infection, vii, 1, 3, 5, 7, 11, 16, 18, 25, 31, 34, 39, 40, 42, 45, 46 infectious disease, 8 infectious diseases, 8 infrared spectroscopy, 245, 246 ingestion, 115 inhibition, 7, 25, 30, 40, 44, 104, 146, 148, 168, 170, 174, 219, 284, 286, 287, 297 inhibitor, 32, 57, 63, 161, 174 inhibitory effect, 30, 46, 104, 106 initiation, 284 injury, 5 inoculation, 233 inoculum, 233, 253, 300 insects, 11, 13 insight, 41, 283 integrity, 15 intensity, 132, 209 interaction, 7, 8, 9, 15, 33, 35, 43, 87, 110, 111, 118, 209, 229, 274 Interaction, 131, 223 interactions, 6, 8, 9, 20, 32, 33, 36, 38, 39, 114, 115, 125, 141, 150, 223, 274, 302 interface, 120, 135, 268 intermolecular interactions, 246 international standards, 251, 253 interpretation, 196, 240 intervention, 102 intestine, 61, 75 intravenously, 62 intrinsic viscosity, 200, 209 invertebrates, viii, 49, 50, 55, 61, 64, 65 ionization, 35, 186, 292 ions, 65, 182, 186, 260, 268, 270 IR, 182, 255 iron, 65, 69, 121, 136, 182 irradiation, 65, 170, 183, 236 Islam, 112, 123 isolation, 23, 24, 33, 45, 85, 141, 185, 291 isophthalic acid, 213 isothermal, 196, 202 isothermal crystallization, 196, 202 isotherms, 108, 145 isotope, 110, 116, 117, 118, 119, 120, 122, 123, 124, 125

311

isotopes, vii, 118, 122 Italy, 51, 52, 74

J Japan, 46, 49, 51, 52, 66, 68, 69, 70, 75, 83, 100, 259 joints, 251

K kidney, 66 kinetic equations, 218 kinetic model, xii, 182, 183, 256, 289 kinetic parameters, 86, 89 kinetics, 68, 107, 109, 110, 111, 113, 118, 123, 137, 144, 145, 146, 147, 150, 172, 182, 202, 261, 285, 294, 301 Korea, 217, 220, 235

L labeling, 62 lactic acid, 221, 235, 260 Lactobacillus, 29, 40, 191, 237 lactones, 192, 208 lactose, 226 lakes, 64 land, 15, 121, 124 landfills, ix, 51, 101, 126 larvae, xii, 70, 235, 274, 277 laser, 271 laws, 150 leachate, 51, 56, 66, 67, 76, 186, 219 leaching, 78, 79, 84, 182 leakage, 7 legislation, 240 lending, 209 lesions, 5 life sciences, 241 ligands, 8 light conditions, 60 lignin, 7, 33, 58, 68, 74, 247, 256 limitation, 93, 110, 218, 230, 246 linkage, 9, 10, 12, 27 lipase, 12, 41, 198, 199, 204, 210 lipases, 12 lipids, 72, 246 liquid chromatography, xii, 72, 186, 290, 301 liquid nitrogen, 241 liquid phase, 90, 230 liquids, 107, 108, 120, 124 Listeria monocytogenes, 29

312

Index

literature, x, 79, 81, 84, 88, 91, 126, 167, 171, 210, 239, 240, 246, 247, 254, 286, 298, 299 liver, 61, 62, 63, 64, 66, 67, 68, 69, 72, 75, 76, 99 location, 97, 107 low density polyethylene, 198, 256 low temperatures, 64 luciferase, 249, 256 luciferin, 160, 249 luminescence, 166 lung, 46 lysine, 230 lysis, 91 lysozyme, 12, 37, 39, 42

M macromolecular chains, 200, 211, 213 macromolecules, 7, 11, 27, 29, 137 macrophages, 44 macropores, 96 magnesium, 143 magnetite, 121 maize, 37, 45, 171 malaria, 162 malt extract, 219 maltose, 219 management, 2, 22, 124, 140 manganese, 58, 68 manipulation, 138 manufacturing, 50, 51, 53, 73, 217, 274 manure, 51, 218, 236 marine environment, 69, 70, 136 market, xii, 190, 218, 235, 240, 254, 289, 290 marrow, 271 mass loss, 111, 198, 203 mass spectrometry, xii, 35, 69, 70, 72, 100, 160, 174, 176, 290, 292, 298, 301, 302 material surface, 241, 242, 248 matrix, 20, 64, 107, 108, 113, 126, 128, 212, 254, 262 maturation, 236, 237 MBP, 64 MDI, 205 measurement, 9, 32, 136, 138 meat, 89, 90, 91 mechanical properties, x, 190, 192, 198, 206, 208, 209, 240, 269 media, 17, 43, 75, 91, 102, 144, 160, 172, 253, 261, 275, 278, 284, 285 median, 109 medicine, 210 Mediterranean, 98

melatonin, xii, 289, 290, 291, 292, 293, 294, 295, 296, 297, 298, 299, 301, 302 melt, 192, 196, 209, 212 melting, x, 189, 190, 195, 196, 198, 199, 200, 201, 202, 203, 204, 205, 206, 209, 210, 211, 212 melting temperature, 196, 203, 204, 209, 210, 212 melts, 211 membrane permeability, 47 membrane separation processes, 235 membranes, 14, 34, 47, 285 men, 63, 302 Merck, 277 metabolic intermediates, 168, 174, 284 metabolic pathways, 131, 160, 171, 224, 298 metabolism, viii, 36, 41, 49, 50, 53, 55, 59, 60, 61, 62, 63, 64, 65, 67, 69, 71, 72, 73, 75, 99, 120, 132, 136, 173, 174, 224, 225, 230, 237, 249, 285, 286, 298, 301, 302 metabolites, viii, xii, 49, 55, 57, 60, 61, 63, 64, 65, 66, 69, 70, 73, 76, 85, 97, 98, 104, 106, 109, 110, 111, 117, 119, 121, 137, 160, 166, 170, 172, 174, 176, 177, 185, 225, 228, 247, 256, 289, 290, 292, 298, 301 metabolizing, 72 metabolomics, 185 methane, 124 methanogenesis, 154, 250 methanol, 179, 181, 237, 247, 291, 292 methionine, 225, 235 methyl group, 178, 182, 246 methyl groups, 178, 246 methylation, 170 methylene, x, 189, 190, 193, 195, 196, 246 methylene group, 190, 196, 246 Mg2+, 58 mice, 45, 46, 64, 69 microbial cells, 242, 294, 295 microbial communities, 250 microbial community, 122, 302 microbiota, viii, 77, 92, 95, 96 microcosms, 69 microenvironment, 268 micrograms, 111 microorganism, xi, 25, 26, 29, 31, 86, 118, 222, 223, 254, 273, 275, 276, 281, 291, 301 microorganisms, vii, viii, xi, 2, 3, 5, 11, 13, 17, 23, 26, 27, 29, 36, 49, 50, 52, 53, 55, 65, 69, 78, 91, 95, 97, 98, 99, 128, 131, 132, 134, 137, 140, 160, 166, 191, 208, 218, 219, 220, 222, 223, 227, 228, 229, 232, 233, 234, 236, 237, 241, 242, 244, 249, 250, 253, 254, 273, 275, 280, 284, 285, 293, 296, 302 microphotographs, 200

Index microscope, 268 microscopy, xi, 239, 242, 243, 270 microsomes, 61, 62, 63, 67, 69, 72, 99 middle lamella, 16 migration, viii, 49, 50, 51, 65, 108, 116, 119, 120 milk, 46, 219, 220, 226 minerals, 121, 145, 261 mining, 112 mixing, 108, 118, 132, 137, 210, 227 MLT, xii, 289, 290, 291, 296, 297, 301 mobility, 84, 102, 201 model system, 269 modeling, 121, 123, 145 models, 96, 107, 108, 112, 122, 123, 129, 130, 144, 149, 150, 263 modulus, 205 moieties, 59, 213 moisture, 126 mole, 223 molecular biology, 34, 36 molecular changes, 247 molecular mechanisms, x, 159 molecular structure, 133, 247 molecular weight, vii, ix, 1, 11, 12, 13, 14, 19, 21, 22, 25, 29, 31, 39, 42, 59, 101, 104, 105, 106, 115, 128, 130, 141, 190, 192, 193, 197, 198, 200, 203, 205, 206, 208, 209, 211, 212, 213, 225, 248 molecular weight distribution, 13, 14, 22, 203, 248 molecules, vii, 1, 2, 3, 5, 9, 10, 12, 17, 20, 22, 24, 25, 28, 30, 128, 130, 132, 140, 160, 179, 246, 274 mollusks, 33 monkeys, 63, 71, 74 monomer, x, 11, 20, 189, 190, 191, 192, 203, 213 monomers, 13, 25, 191, 192, 197, 200, 208, 213 mononuclear cells, 25 monosaccharide, 9, 27 morbidity, 260 morphogenesis, 32 morphology, 98, 244, 254, 261 mRNA, 62, 64, 268 mucoid, 25, 27, 35, 40 mutant, 37 Mycobacterium, 130

N NaCl, 89, 90, 91, 219, 222, 276, 277, 285 NAD, 57 NADH, 57 naphthalene, 110 National Research Council, 108, 122 natural environment, 78, 89, 140, 160, 241 natural gas, 125

313

necrosis, 41, 44 neglect, 109, 115 network, 15, 32, 42 neurotoxic effect, 81 nitrate, vii, 65, 85, 91, 106, 125, 135, 139, 142, 143, 182, 186, 236 nitric oxide, 44 nitrification, 170 nitrogen, vii, 84, 85, 87, 89, 91, 92, 95, 139, 143, 168, 170, 171, 178, 220, 222, 235, 292, 300 nitrogen gas, 222, 292 nitroso compounds, 104 NMR, x, xi, 159, 172, 173, 174, 175, 176, 177, 178, 182, 184, 185, 197, 202, 239, 247, 255 novelty, 97 nuclear magnetic resonance, 46 nucleation, 202 nuclei, 178 nucleotide sequence, 40 nutrients, ix, 58, 78, 89, 90, 91, 92, 110, 128, 130, 133, 143, 152, 225, 230, 233

O observations, 107, 119, 196, 200, 205, 211, 240, 254, 261, 269, 284 oedema, 68 oil, 104, 112, 119, 121, 123, 124, 140, 157, 158, 191, 217, 286 oil refineries, 112 oil spill, 104, 119, 123, 140 oils, 118, 217 oligomers, 5, 11, 16, 25, 28, 30, 37, 38, 41, 42, 44, 59, 191, 192, 193, 197, 200, 206, 208 oligosaccharide, vii, 1, 2, 3, 5, 8, 9, 13, 29, 30, 37, 39, 43, 45 oligosaccharides, vii, 1, 2, 3, 4, 5, 6, 7, 8, 9, 10, 11, 12, 13, 14, 16, 19, 20, 21, 22, 24, 26, 27, 28, 29, 30, 31, 33, 35, 36, 37, 38, 39, 40, 41, 42, 44, 45, 46, 47 optical density, 293 optical microscopy, 241, 242 optimization, 84, 121, 230, 277 organ, 61 organic chemicals, 70, 115, 116, 122 organic compounds, vii, 55, 91, 92, 102, 103, 105, 112, 115, 121, 144, 225, 230, 293 organic growth, 85 organic matter, vii, 65, 67, 79, 83, 104, 105, 132, 133, 142, 144, 170, 179, 218, 224, 225, 226, 230, 238, 251 organic solvent, 81, 182, 247, 300 organic solvents, 81

314

Index

organism, xi, 6, 87, 97, 103, 273, 282 organophosphates, vii osteoclasts, 264, 268, 270 osteocytes, 262 osteotomy, 269 overload, 89 overproduction, 248 oxidation, 57, 58, 61, 65, 81, 100, 105, 131, 135, 168, 180, 182, 183, 221, 230, 234, 245, 250, 275, 287, 300 oxidation products, 61, 100, 245, 250 oxygen, 32, 41, 58, 65, 89, 91, 105, 112, 134, 135, 142, 143, 144, 179, 220, 222, 223, 224, 229, 230, 232, 237, 245, 250, 252, 253, 255 Oxygen, 134, 153, 157, 158, 168 oxygen consumption, 224 oxygen consumption rate, 224 ozonation, 180, 185, 275, 285 ozone, 180

P packaging, 50 Pakistan, 170, 236 parameter, 109, 117, 125, 149, 209, 213, 225, 230, 234, 277 parameter estimation, 117 parasite, 38 particles, ix, 127, 128, 132, 139, 145, 152, 242, 247 particulate matter, 145 partition, ix, 50, 101, 103, 104, 149 passive, 102, 252 pathogenesis, 6, 33, 46 pathogens, vii, 1, 2, 6, 8, 11, 16, 17, 27, 31, 34, 36, 37, 40, 41, 46 pathology, 38 pathways, viii, x, 32, 49, 55, 62, 64, 66, 69, 99, 119, 131, 159, 160, 171, 175, 176, 224 PCR, 220, 222 pectin, viii, 2, 16, 17, 22, 34, 41, 45 penicillin, xi, 273 perception, 5, 27 performance, xii, 34, 46, 89, 92, 93, 142, 190, 222, 271, 290, 301 perfusion, 69 permeability, 7, 105, 107, 108, 109, 113, 119, 141, 240 permeation, 8, 13, 32 permit, 140 peroxide, 143, 296 personal, x, 66, 239, 277, 292 perylene, 130 pesticide, 41, 78, 84, 96, 97, 100, 144, 171, 186

pesticides, viii, 2, 77, 78, 79, 82, 96, 98, 99, 161, 178 pests, 2 PET, 190, 192, 198, 208 petrochemicals, xi, 273 petroleum products, vii pH, xii, 52, 57, 58, 60, 81, 83, 89, 90, 91, 93, 95, 103, 133, 139, 168, 182, 198, 199, 204, 210, 218, 219, 220, 223, 224, 229, 232, 233, 234, 276, 286, 289, 290, 291, 292, 294, 295, 296, 297, 298, 301 phase transformation, 270 PHB, 237 phenol, viii, xi, xii, 49, 50, 53, 59, 63, 64, 65, 182, 249, 273, 274, 275, 276, 277, 278, 280, 281, 282, 283, 284, 285, 286, 287 phenolic resins, 51 phenotypes, 262 phenylalanine, 3, 37 Phenylalanine, 44, 231, 234 phosphate, vii, xi, 143, 198, 199, 204, 259, 260, 261, 262, 263, 264, 265, 266, 267, 268, 269, 270, 271 phosphates, 170, 260, 268, 269 phosphorous, 228 phosphorus, 260 phosphorylation, 30, 36 photodegradation, x, 65, 79, 159, 160, 178, 179, 182, 183 photographs, 202 photolysis, 65, 81, 160, 178, 179, 180, 182, 186 photons, 179, 249 photooxidation, 65, 182 physical properties, ix, 127, 144 physical treatments, 128 physicochemical methods, 275, 285 physicochemical properties, 79, 81, 82, 92 physico-chemical properties, 170 physics, 240 physiology, 185 phytopathogen, vii, 1, 2, 3, 5, 6, 25, 29, 30, 31 phytoplankton, 65 phytoremediation, 122 pineal gland, xii, 289, 290 placenta, 66 plankton, viii, 49 plants, viii, x, xii, 2, 3, 5, 6, 7, 9, 10, 13, 15, 16, 21, 22, 25, 26, 27, 30, 31, 35, 36, 38, 40, 42, 43, 44, 45, 46, 49, 50, 51, 53, 55, 60, 65, 67, 71, 79, 99, 112, 159, 160, 162, 163, 164, 167, 171, 178, 183, 185, 219, 225, 235, 237, 274, 289, 290, 302 plaque, 244 plasma, 10, 33, 36, 38, 43, 62 plasma membrane, 10, 33, 36, 38, 43 Plasmodium falciparum, 185 plastics, 50, 51, 255

Index PM, 71, 75 polarization, 247 pollutants, ix, x, xi, 60, 94, 101, 104, 106, 107, 109, 110, 111, 112, 113, 116, 119, 122, 140, 159, 160, 164, 183, 185, 273, 274, 284, 286, 290 pollution, ix, 51, 69, 101, 102, 103, 106, 108, 111, 114, 115, 116, 121, 167, 192, 236, 274, 275, 286 poly(ethylene terephthalate), 208 polycarbonate, viii, 49, 50, 70, 72 polycarbonate plastics, viii, 49, 50 polycarbonates, 50 polychlorinated biphenyls (PCBs), 104 polycondensation, 192, 193, 200, 202, 206, 208, 213 polycyclic aromatic hydrocarbon, 120, 124, 158 polydispersity, 248 polyesters, x, 50, 189, 190, 191, 192, 193, 194, 195, 196, 197, 198, 199, 200, 202, 203, 205, 208, 210, 211, 212, 213 polyethylene, 243, 245, 248, 252, 254, 255, 256, 257 polyethylenes, 254, 255 polymer, vii, x, 1, 9, 11, 13, 16, 18, 190, 196, 198, 208, 209, 210, 212, 239, 240, 241, 245, 247, 248, 251, 254, 256, 260 polymer blends, 209, 210 polymer composites, 241, 256 polymer film, 256 polymer films, 256 polymer materials, xi, 239, 245, 248, 251 polymer molecule, 245 polymer nanocomposites, 247 polymeric chains, 245 polymeric materials, 240, 245, 260 polymerization, 11, 13, 18, 28, 29, 31, 59, 61, 72, 192, 193, 202, 208 polymerization process, 192 polymers, x, 2, 11, 16, 20, 36, 53, 61, 162, 189, 190, 191, 193, 194, 196, 197, 203, 206, 208, 209, 210, 212, 213, 241, 245, 246, 247, 256, 260 polysaccharide, vii, 1, 3, 6, 11, 15, 23, 25, 27, 30, 36, 37, 42, 43, 44, 249 polyvinyl chloride, 51, 75 pools, 119, 124 poor, ix, 13, 58, 84, 101, 104, 130, 134, 246, 248 population, 66, 75, 122, 128, 132, 133, 138, 139, 217, 224, 278 porosity, 95, 119, 139, 144, 146, 149, 152, 212 porous media, 109, 119, 126 Portugal, viii, 77, 82, 273, 276 potassium, 85 potato, 28, 61, 75, 236 power, 2 precipitation, 45, 72, 80, 182, 269, 270 prediction, 107, 108, 295

315

preference, 27 pregnancy, 62, 63, 71 pressure, 13, 82, 133, 191, 240 prevention, 102, 237 probability, 150, 200 probability distribution, 150 process control, 230 producers, 254 production, vii, viii, x, xi, xii, 1, 2, 3, 6, 12, 13, 14, 17, 20, 21, 25, 27, 31, 36, 37, 38, 39, 41, 42, 44, 47, 49, 50, 51, 53, 60, 104, 112, 125, 161, 162, 163, 167, 185, 186, 189, 190, 191, 192, 213, 217, 219, 220, 222, 223, 224, 227, 229, 230, 233, 234, 236, 237, 239, 240, 250, 251, 252, 253, 289, 290 productivity, 14, 235 program, 153, 243 proliferation, 73, 267, 271 promote, 79, 94, 96, 223 promoter, 41 propagation, 243 propane, viii, 49, 50, 53 propylene, x, 189, 190, 191, 193, 194, 196, 197, 198, 199, 200, 201, 202, 204, 205, 206, 207, 208, 209, 211, 213 protease inhibitors, 3 proteases, 12 protective coating, 50 protein, 9, 10, 30, 33, 35, 36, 46, 61, 62, 217, 235, 246, 248, 249, 255, 256 proteinase, 30, 32 proteins, 6, 7, 8, 33, 43, 46, 72, 226, 235, 246, 248, 249, 260 protocol, 243, 248 protocols, x, 239 protons, 193 prototype, 260 protozoa, 131 Pseudomonas aeruginosa, 25, 26, 27, 35, 40, 42, 300, 302 public health, viii, 77, 166 pulse, 168, 169 pulses, 168, 169 pure water, 65, 102, 180 purification, 35, 41, 47, 55, 172, 176 PVA, 236 PVC, 50, 51 pyrene, 126, 157

Q quantitative estimation, 163 quantum yields, 179 quaternary ammonium, 39

316

Index

quinone, 63, 64

R race, 44, 123 radial distance, 152 radiation, 64, 70, 186 Radiation, 236 radio, 160 radius, 146, 147 rain, 82 range, ix, 11, 12, 13, 50, 51, 52, 57, 64, 70, 78, 83, 89, 91, 133, 147, 148, 173, 178, 183, 196, 209, 229, 240, 275, 285, 290, 292, 295, 301 reactive oxygen, 2, 5, 40, 60, 65, 72, 76 reactivity, 53, 130 reagents, 248, 291 reasoning, 114 receptors, 5, 36, 102, 105, 106, 112, 114 recognition, 5, 20, 44, 47 recovery, 36, 65, 69, 118, 138, 190, 217, 235, 275 recrystallization, 196 recycling, 51, 53, 67, 72, 218 reduction, 71, 78, 89, 90, 95, 106, 109, 112, 116, 121, 123, 135, 142, 221, 224, 225, 230, 234, 237, 283, 284 refining, xi, 273, 284 reflection, 246 refractory, 104 regeneration, ix, xi, 78, 93, 94, 95, 96, 98, 100, 259, 260, 262, 268 Registry, viii, 49, 50 regression, 158 regulation, 67, 230 regulations, 128, 217 reinforcement, 2, 5, 27 relationship, 30, 74, 86, 294, 295, 297 relationships, 24, 31, 34 relaxation, 197 relevance, xi, 239 reliability, 109, 145 remediation, ix, xii, 102, 105, 107, 108, 116, 117, 118, 119, 120, 121, 122, 124, 125, 126, 140, 141, 274 remodeling, 268 repair, 262, 268, 270 repression, 104, 287 reproduction, xii, 68, 289, 290 residues, 2, 3, 9, 10, 11, 12, 15, 16, 20, 22, 23, 24, 27, 28, 29, 30, 37, 78, 97, 100, 156, 170, 171, 186, 217, 237 resilience, 190 resins, 50, 51

resistance, viii, 2, 3, 5, 6, 7, 31, 32, 35, 37, 38, 40, 41, 46, 53, 58 resolution, 209, 270 resorcinol, xii, 274, 275, 276, 277, 278, 279, 280, 281, 282, 283, 284, 285, 286 resources, x, 124, 128, 130, 189, 190, 213 respiration, 137, 184 restitution, 95 retardation, 120, 145 retention, 91, 277, 298 retina, 290 RF, 43 rheological properties, 240 rice, viii, ix, 7, 28, 35, 37, 43, 46, 47, 77, 78, 79, 82, 84, 89, 92, 93, 94, 96, 97, 98, 99, 100 rice field, viii, ix, 77, 78, 79, 82, 89, 92, 98, 99, 100 rings, 10, 50, 53, 104, 128, 130, 131 risk, ix, 101, 102, 103, 104, 105, 106, 107, 111, 113, 114, 115, 116, 122, 125, 130, 150, 166, 187, 252 risk assessment, 122, 125, 187 RNA, 7 rodents, 63 rods, 222 room temperature, 130, 204, 205, 208, 209, 210, 212, 221 Royal Society, 185 rubber, 162, 163, 166, 167, 168

S SA, 32, 47, 75, 276, 291 safety, 243 salinity, xi, xii, 89, 91, 103, 133, 273, 275, 276, 278, 285, 289, 290, 292, 297, 298, 301 salmon, 64, 68 Salmonella, 29 salt, xi, 23, 24, 31, 208, 219, 222, 259, 275, 276, 277, 285, 291 salts, xii, 76, 136, 167, 182, 218, 274 sample, 51, 52, 94, 192, 197, 205, 213, 219, 221, 225, 231, 234, 237, 241, 242, 244, 246, 247, 248, 249, 250, 251, 252, 254, 278, 292 sampling, ix, 83, 102, 107, 108, 109, 138, 252, 283 saturation, 294 scanning electron microscopy, 244 science, 218, 240 scientific community, 240 SCP, 237 seafood, 218, 235 search, 220, 235, 240, 254 searches, 292 searching, 2 SEC, 208

Index secrete, 12, 132 secretion, 302 security, 218 sediment, ix, xii, 51, 52, 57, 72, 73, 74, 75, 127, 128, 129, 130, 131, 132, 134, 135, 136, 137, 138, 139, 140, 141, 142, 144, 145, 148, 149, 150, 152, 156, 289, 290, 291, 300, 302 sediments, ix, 50, 52, 53, 66, 68, 73, 75, 76, 105, 117, 127, 128, 130, 131, 132, 133, 134, 135, 136, 137, 138, 139, 140, 142, 143, 144, 145, 146, 148, 150, 152, 153, 157, 158 seed, 37, 219, 221, 224, 236, 237 seedlings, 32, 82 SEM micrographs, 200, 202, 210 semen, xii, 289, 290, 302 sensitivity, 119, 125, 246, 247, 248, 249, 278, 282, 283 separation, xi, 37, 39, 110, 217, 239, 247, 275, 277, 292, 298 septum, 252, 253 sequencing, 24, 37 series, 5, 19, 21, 193, 194, 200, 211, 212, 278, 284, 292 serum, 36, 62, 74, 137, 261, 270 serum albumin, 270 sewage, 51, 56, 66, 67, 69, 70, 236 sex, 73 sex hormones, 73 sexual activity, xii, 289, 290 shape, 28, 43 shrimp, 46, 238, 277 sign, 246 signal transduction, 30, 32, 34, 37 signaling, 3, 5, 36, 38, 43, 44 signals, 2, 6, 30, 32, 36, 43, 44, 47, 240, 245, 246 silica, 292 similarity, 10, 45, 86, 220, 223, 293 simulation, 70, 119, 193 sites, ix, 9, 10, 34, 50, 51, 66, 74, 75, 78, 82, 84, 102, 105, 108, 120, 121, 123, 124, 127, 128, 138, 140, 141, 145, 146, 149, 152, 219, 263, 271 size-exclusion chromatography, 34 skin, 186 sludge, x, 53, 55, 66, 70, 159, 160, 163, 167, 168, 170, 184, 185, 224, 236, 238, 300, 301 sodium, xii, 34, 52, 72, 208, 238, 242, 248, 274, 275, 276, 278, 279, 281, 284 sodium hydroxide, 248 soil, vii, ix, 26, 46, 50, 51, 55, 56, 57, 72, 75, 78, 79, 82, 83, 84, 85, 99, 100, 101, 102, 103, 104, 107, 108, 109, 113, 114, 115, 116, 118, 119, 120, 121, 122, 123, 124, 126, 128, 143, 144, 145, 160, 166,

317

170, 171, 177, 183, 185, 186, 208, 219, 224, 230, 233, 250, 251, 253, 255, 256, 284, 285 soil particles, 78, 115 soil pollution, ix, 101, 102, 103, 116 solid matrix, 95 solid phase, 96, 146, 149, 152, 240, 247 solid state, 19, 256 solid waste, 56, 74, 76 solubility, xi, 11, 50, 79, 81, 84, 97, 108, 120, 128, 130, 133, 136, 140, 148, 212, 230, 259, 260 solvent, 117, 121, 125, 140, 179, 182, 209, 210, 275, 277 solvents, ix, 11, 101, 105, 107, 108, 117, 119, 120, 121, 125, 126, 207 sorption, 102, 103, 106, 107, 108, 109, 128, 132, 133, 144, 145, 146, 147, 148, 149, 150, 151, 152 sorption isotherms, 146 sorption kinetics, 144, 145, 146 sorption process, 108 soybean, 3, 9, 10, 22, 26, 28, 30, 31, 33, 34, 38, 40, 44, 46, 47, 59, 66, 237 speciation, 236 species, 2, 5, 6, 15, 17, 18, 20, 34, 40, 41, 45, 46, 50, 59, 60, 62, 63, 65, 72, 76, 85, 86, 87, 110, 128, 130, 140, 222, 223, 237, 263, 275, 281 specificity, viii, 2, 11, 17, 20, 26, 27, 31, 35, 40, 42, 45, 46, 171, 182, 284 spectrophotometry, 169 spectroscopy, 46, 197, 245, 247, 255, 256, 292 spectrum, 3, 178, 240, 247, 298 speed, 13, 192 sperm, 98, 302 spine, 269 spore, 38, 235 Sprague-Dawley rats, 63 stability, 81, 89, 198, 203 stabilization, 225, 230, 238 stages, 20, 64, 69, 191, 192, 197, 203, 255, 269 standard deviation, 109, 166, 293, 296 standards, ix, 101, 103, 105, 113, 114, 115, 123, 125, 253, 254, 277, 292, 301 Staphylococcus, 7, 29 starch, 219, 253, 260 statistical analysis, 262 statistics, 32 sterile, 219, 220, 221, 276 stochastic model, 150 stock, 182, 191 stoichiometry, 143, 148 storage, ix, 2, 101, 102, 120, 138, 233, 235, 238 strain, xii, 12, 16, 21, 23, 24, 26, 27, 35, 36, 37, 38, 41, 42, 43, 44, 47, 55, 56, 57, 58, 59, 63, 69, 70, 71, 72, 73, 76, 85, 87, 97, 98, 99, 171, 172, 174,

318

Index

176, 177, 185, 222, 256, 274, 276, 277, 281, 285, 286, 289, 291, 293, 300 strategies, 14, 78, 97, 107, 108, 113, 129, 132, 136, 140, 144, 156, 170, 236, 276 strength, 182, 198, 213, 227, 275, 279 stress, 60, 138, 177, 190 stretching, 245, 246 stromal cells, 268, 271 structural changes, 247 structure formation, 270 substitutes, 269, 271 substitution, 19, 22, 37, 86, 300 substrates, vii, 13, 26, 27, 31, 36, 61, 104, 132, 133, 137, 228, 275, 282, 286, 287 sugar, 2, 8, 13, 218 sugar beet, 218 sulfate, vii, 20, 21, 46, 61, 62, 63, 64, 82, 85, 134, 135, 139, 142, 143, 155, 237, 242, 300, 301 sulfur, vii, 21, 22, 81, 87, 184, 225 summer, 61, 83 Superfund, 117 supply, 102, 105, 135, 142, 220, 229, 230, 232, 233, 250, 252 suppression, 31, 73 surface area, 95, 139, 246 surface layer, 131, 134, 136, 240, 246 surface properties, 98 surfactants, 140 susceptibility, 12, 31, 40, 42 suspensions, 79 Sweden, 115 swelling, 240 symptom, 40 symptoms, 186 synthesis, x, 3, 5, 10, 34, 38, 40, 44, 46, 67, 162, 184, 189, 190, 191, 192, 204, 205, 206, 207, 213, 302 synthetic polymeric materials, 249 systemic circulation, 63 systems, 13, 30, 44, 60, 78, 84, 89, 91, 95, 108, 110, 116, 120, 142, 143, 163, 165, 167, 168, 171, 183, 184, 208, 230, 274, 276

T tandem mass spectrometry, 186 tanks, 51, 120 tanning, xi, 273 tannins, 274 tar, 119 target organs, 185 targets, 102, 106, 112 taxonomy, 32

TCP, xi, 259, 260, 261, 262, 263, 264, 265, 268, 269 technology, 95, 131, 140, 141, 142, 210, 213, 275 teeth, xi, 259, 260 temperature, xii, 52, 57, 58, 60, 61, 64, 68, 70, 79, 82, 89, 91, 95, 103, 110, 132, 139, 171, 191, 192, 193, 196, 197, 198, 209, 210, 212, 222, 241, 248, 251, 269, 289, 290, 291, 292, 297, 298, 300, 301 tensile strength, 198, 203, 240 terpenes, 274 testosterone, xii, 62, 289, 290, 302 tetrachloroethylene, 115 TGA, 197, 204, 205 theory, 125, 209 thermal decomposition, 203 thermal degradation, 197, 203 thermal properties, 190 thermal stability, 198, 204, 205 thermograms, 198, 203, 204 thermogravimetric, 197 thermophil, 19 thermostability, 197, 198 Thessaloniki, 189 thin films, 210 threat, 106, 128, 130 threonine, 235 threshold, 111 thresholds, 78 thymine, 130 tibia, 270 time, ix, xi, xii, 2, 16, 20, 52, 58, 63, 82, 84, 85, 88, 91, 101, 102, 103, 104, 106, 109, 110, 112, 113, 114, 120, 137, 138, 139, 140, 147, 149, 150, 152, 165, 166, 168, 172, 179, 180, 192, 197, 199, 201, 204, 208, 220, 230, 232, 233, 234, 240, 242, 243, 250, 251, 253, 254, 273, 277, 279, 282, 283, 284, 289, 290, 294, 298 time frame, ix, 101, 102, 103, 106, 208 timing, 123 tissue, 16, 41, 45, 62, 260, 261, 263, 264, 267, 268, 271, 302 titanium, 192, 271 tobacco, 10, 28, 33, 39, 40, 41, 44, 45, 71 tofu, 42 toluene, vii, 106, 110, 125, 133, 179 total product, 162 toxic effect, 53, 58, 66, 160, 225 toxic substances, 78, 92, 274 toxicity, viii, x, xi, xii, 2, 49, 55, 63, 64, 65, 66, 69, 76, 78, 81, 98, 99, 102, 104, 106, 111, 115, 125, 130, 143, 159, 160, 164, 166, 225, 237, 273, 274, 275, 276, 277, 282, 283, 285, 286, 287, 296 toxicology, 100 transcription, 7

Index transesterification, 202 transference, 281 transformation, 65, 78, 81, 84, 86, 87, 119, 125, 149, 150, 171, 175, 177, 179, 183, 185, 261, 286, 299 transformation product, 78, 84 transformations, 184, 186 transforming growth factor, 269 transgenic, 35, 40 transition, 209, 269 transition temperature, 209 translocation, 71 transmission, 241, 242, 245, 246 transport, ix, 50, 101, 102, 106, 108, 110, 118, 120, 121, 124, 128, 129, 132, 141, 146, 150, 152, 218 transport processes, 120, 132, 150, 152 treatment methods, 137, 275 trend, 83, 200, 224, 229 trial, 229, 230, 231 trichloroethylene, 115 trimer, 29 turnover, 142 Tyrosine, 231, 234

U U.S. Geological Survey, 125 ultrasound, 242 uncertainty, ix, 101, 107, 108, 109, 111, 112, 113, 114, 119, 125, 131, 155 uniform, 124 urea, 170, 171, 178, 182 urethane, 205, 213 urine, 63, 70, 71 UV, 81, 86, 87, 90, 99, 169, 172, 173, 182, 183, 277

V vacuum, 192 validation, 107 values, viii, 50, 52, 77, 79, 83, 84, 86, 89, 91, 95, 120, 149, 166, 193, 196, 198, 205, 209, 210, 223, 225, 229, 230, 232, 250, 283 variability, 108, 109, 118, 120, 248 variable, 83, 105, 106, 112, 123, 126, 274 variables, 68, 145, 147, 152, 158 variance, 107, 109, 110, 112, 115 variation, 107, 108, 109, 110, 171, 195, 209 vegetables, 60, 70, 74, 75 vegetation, 79 versatility, xi, 273 vertebrates, viii, 49, 50, 55, 65, 260 vibration, 245, 246

319

vinasse, 235 vinylchloride, 105, 111, 117 viral infection, 60 virus, 6, 28, 41, 44, 45 viruses, 39 viscosity, 11, 13, 14, 25 visualization, 119, 243 vitamins, 85 volatility, 79 volatilization, 78, 79, 84, 100, 102, 103, 104, 109, 115 vulcanization, 162, 164 vulnerability, 89

W waste disposal, x, 102, 239 waste disposal sites, 102 waste water, 186 wastewater, viii, x, xi, 49, 50, 51, 52, 55, 65, 66, 67, 70, 78, 92, 159, 160, 162, 163, 164, 165, 167, 171, 178, 183, 184, 185, 186, 217, 220, 221, 226, 228, 230, 232, 233, 234, 235, 236, 237, 273, 274, 285, 286 wastewater treatment, viii, x, 49, 50, 51, 52, 55, 65, 78, 92, 159, 160, 162, 163, 164, 165, 167, 171, 178, 183, 185, 217, 230, 237, 285 wastewaters, xi, 51, 163, 164, 170, 180, 186, 217, 218, 220, 221, 226, 232, 233, 235, 237, 273, 286, 287 water absorption, 204, 207 water policy, 275, 286 water quality, 116 wavelengths, 182, 183 wealth, 241 weight loss, 197, 198, 199, 200, 203, 204, 211 weight ratio, 209 weight reduction, 198 wells, 84 wetlands, 105 wheat, 19, 36, 40, 41, 46, 236, 247, 256 wheat germ, 41 wildlife, 53, 67 windows, 124 winter, 61, 67, 83, 161 women, 63, 74 wood, 19, 192 workers, 166, 186

320

Index

X xanthan, viii, 2, 5, 22, 23, 24, 25, 31, 33, 34, 36, 37, 40, 42, 43, 45 xenobiotics, x, xii, 61, 78, 96, 126, 159, 160, 184, 247, 255, 274, 276, 284 X-ray analysis, 41 XRD, 203 xylan, viii, 2, 18, 19, 20, 42 xylene, vii, 123 xylenes, 106, 110 xyloglucan, vii, 1, 3, 5, 15, 16, 20, 35, 36, 37, 40, 42, 44, 46, 47

Y yeast, xii, 53, 67, 74, 89, 91, 219, 276, 281, 286, 287, 289, 290, 292, 296, 300 yield, 12, 16, 21, 39, 89, 90, 135, 146, 152, 168, 179, 218, 235, 236 yolk, 64, 68

Z zebrafish, 61, 62, 71 zooplankton, 65