Environmental degradation and transformation of organic chemicals

Environmental Degradation and Transformation of Organic Chemicals

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Environmental Degradation and Transformation of Organic Chemicals Alasdair H. Neilson and Ann-Sofie Allard

Boca Raton London New York

CRC Press is an imprint of the Taylor & Francis Group, an informa business

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CRC Press Taylor & Francis Group 6000 Broken Sound Parkway NW, Suite 300 Boca Raton, FL 33487-2742 © 2008 by Taylor & Francis Group, LLC CRC Press is an imprint of Taylor & Francis Group, an Informa business No claim to original U.S. Government works Printed in the United States of America on acid-free paper 10 9 8 7 6 5 4 3 2 1 International Standard Book Number-10: 0-8493-7241-0 (Hardcover) International Standard Book Number-13: 978-0-8493-7241-4 (Hardcover) This book contains information obtained from authentic and highly regarded sources. Reprinted material is quoted with permission, and sources are indicated. A wide variety of references are listed. Reasonable efforts have been made to publish reliable data and information, but the author and the publisher cannot assume responsibility for the validity of all materials or for the consequences of their use. No part of this book may be reprinted, reproduced, transmitted, or utilized in any form by any electronic, mechanical, or other means, now known or hereafter invented, including photocopying, microfilming, and recording, or in any information storage or retrieval system, without written permission from the publishers. For permission to photocopy or use material electronically from this work, please access www.copyright.com (http:// www.copyright.com/) or contact the Copyright Clearance Center, Inc. (CCC) 222 Rosewood Drive, Danvers, MA 01923, 978-750-8400. CCC is a not-for-profit organization that provides licenses and registration for a variety of users. For organizations that have been granted a photocopy license by the CCC, a separate system of payment has been arranged. Trademark Notice: Product or corporate names may be trademarks or registered trademarks, and are used only for identification and explanation without intent to infringe. Library of Congress Cataloging-in-Publication Data Neilson, Alasdair H. Environmental degradation and transformation of organic chemicals / Alasdair H. Neilson and Ann-Sofie Allard. p. cm. Rev. ed. of: Organic chemicals. c2000. Includes bibliographical references and index. ISBN 978-0-8493-7241-4 (alk. paper) 1. Aquatic organisms--Effect of water pollution on. 2. Organic water pollutants--Environmental aspects. I. Allard, Ann-Sofie. II. Neilson, Alasdair H. Organic chemicals. III. Title. QH545.W3N45 2008 577’.14--dc22

2007012219

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Contents Preface ............................................................................................................................................xix Acknowledgments ........................................................................................................................ xxiii Authors ...........................................................................................................................................xxv

SECTION I Degradation and Transformation Processes Chapter 1

Abiotic Reactions .........................................................................................................3

Introduction........................................................................................................................................3 Photochemical Reactions in Aqueous and Terrestrial Environments ...............................................3 The Diversity of Photochemical Transformations .............................................................................4 The Role of Hydroxyl Radicals ..............................................................................................4 Illustrative Examples of Photochemical Transformations in Aqueous Solutions ..................5 Hydroxyl Radicals in the Destruction of Contaminants ....................................................................9 Fenton’s Reagent .....................................................................................................................9 Polyoxometalates (Heteropolyacids) ..................................................................................... 10 Photolytic Degradation on TiO2 ........................................................................................... 10 Other Photochemically Induced Reactions...................................................................................... 11 The Role of Humic Matter: Singlet Dioxygen ................................................................................. 12 Interactions between Photochemical and Other Reactions ............................................................. 13 Reactions in the Troposphere........................................................................................................... 14 Survey of Reactions .............................................................................................................. 15 Survey of Reactants .............................................................................................................. 17 Aliphatic Hydrocarbons ............................................................................................. 17 Aromatic Hydrocarbons............................................................................................. 17 Biogenic Terpenes ...................................................................................................... 17 Reentry of Tropospheric Transformation Products ......................................................................... 18 Halogenated Alkanes and Alkenes....................................................................................... 18 Arenes and Nitroarenes ........................................................................................................20 Alkylated Arenes .................................................................................................................. 21 Sulfides and Disulfides ......................................................................................................... 21 Chemically Mediated Transformations ........................................................................................... 22 Hydrolysis ............................................................................................................................. 22 Reductive Displacement: Dehalogenation and Desulfurization ...........................................25 Reductive Dehalogenation ...............................................................................................................25 Zero-Valent Metals ...............................................................................................................25 Chromous Chloride ..............................................................................................................26 Porphyrins and Corrins ........................................................................................................ 26 Thiol Reductants .............................................................................................................................. 27 Reductive Desulfurization ...............................................................................................................28 Reductions Other than Dehalogenation ................................................................................28 Nucleophilic Reactions ......................................................................................................... 29

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Oxidations ............................................................................................................................. 30 Anodic Oxidation....................................................................................................... 30 Ozone ...................................................................................................................... 30 Hydrogen Peroxide..................................................................................................... 32 Manganese Dioxide ................................................................................................... 32 Sulfate Radicals ......................................................................................................... 32 Halogenation ......................................................................................................................... 32 Thermal Reactions ................................................................................................................34 Gas Phase ...................................................................................................................34 In Water...................................................................................................................... 36 Electrocatalytic Reactions .................................................................................................... 37 References ........................................................................................................................................ 38 Chapter 2

Biotic Reactions.......................................................................................................... 51

Part 1 Microbial Reactions .............................................................................................................. 51 Introduction .......................................................................................................................... 51 Definitions—Degradation and Transformation .................................................................... 51 Biodegradation of Enantiomers ............................................................................................ 54 Sequential Microbial and Chemical Reactions .................................................................... 55 The Spectrum of Organisms................................................................................................. 57 Bacteria in Their Natural Habitats............................................................................. 57 Marine and Oligotrophic Bacteria ............................................................................. 58 Lithotrophic Bacteria .................................................................................................60 Phototrophic Organisms ............................................................................................ 61 Aerobic and Facultatively Anaerobic Bacteria .......................................................... 62 Bacterial Metabolism of C1 Compounds: Methanotrophs, Methylotrophs, and Related Organisms .............................................................................................. 69 Methane Monooxygenase and Related Systems ........................................................ 69 Anaerobic Bacteria ............................................................................................................... 72 Clostridia.................................................................................................................... 73 Anaerobic Sulfate-Reducing Bacteria ....................................................................... 73 Other Anaerobic Bacteria .......................................................................................... 73 Organisms from Extreme Environments .............................................................................. 74 Psychrophiles ............................................................................................................. 74 Thermophiles ............................................................................................................. 75 Alkaliphiles................................................................................................................ 75 Halophiles .................................................................................................................. 75 Eukaryotic Microorganisms: Fungi and Yeasts ................................................................... 75 Metabolism by Fungi ................................................................................................. 75 Metabolism by Yeasts ................................................................................................ 77 References ............................................................................................................................. 78 Part 2 Reactions Mediated by Higher Organisms ........................................................................... 91 Introduction .......................................................................................................................... 91 Metabolism by Fish .............................................................................................................. 93 Metabolism by Other Organisms.......................................................................................... 95 Mussels ...................................................................................................................... 95 Insects ...................................................................................................................................96 Invertebrates .........................................................................................................................96 Isopods ...................................................................................................................... 97 Oligochaetes ...............................................................................................................97 Polychaetes .................................................................................................................97

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Higher Plants ........................................................................................................................97 Other Organisms...................................................................................................................99 References ........................................................................................................................... 100 Chapter 3

Mechanistic Aspects ................................................................................................ 103

Part 1 Introduction of Oxygen ....................................................................................................... 103 Monooxygenation ............................................................................................................... 103 Hydroxylation of Alkanes ........................................................................................ 103 Epoxidation of Alkenes ........................................................................................... 104 Cycloalkanone Monooxygenases ............................................................................. 105 Monooxygenation of Aromatic Compounds ............................................................ 106 Reactions of Aromatic Hydrocarbons ..................................................................... 107 Reactions of Hydroxybenzoates and Related Compounds ...................................... 108 Monooxygenation of Phenols ................................................................................... 110 Cytochrome P450 Systems ................................................................................................. 113 Prokaryotic Organisms ............................................................................................ 114 Eukaryotic Organisms ............................................................................................. 116 Dioxygenation ..................................................................................................................... 118 Dioxygenases Involved in Dihydroxylation of Arenes ............................................ 118 Ring-Fission Dioxygenases ...................................................................................... 123 Incorporation of Oxygen from Water: Oxidoreductases and Hydratases........................... 129 Oxidoreductases ....................................................................................................... 129 Hydratases ................................................................................................................ 130 Oxidases, Peroxidases, and Haloperoxidases ..................................................................... 131 Oxidases ................................................................................................................... 131 Peroxidases .............................................................................................................. 133 Haloperoxidases ....................................................................................................... 134 References ........................................................................................................................... 135 Part 2 Electron Acceptors Other than Oxygen .............................................................................. 147 Introduction ........................................................................................................................ 147 Oxyanions: Nitrate, Sulfate, Chlorate, Selenate, and Arsenate .......................................... 148 Nitrate and Related Compounds .............................................................................. 148 Sulfate and Related Compounds .............................................................................. 150 Chlorate and Perchlorate.......................................................................................... 150 Selenate and Arsenate .............................................................................................. 151 Metal Cations and Oxyanions ............................................................................................ 152 V(V), Mn(IV), Fe(III), Tc(VII), and U(VI) ............................................................. 152 Humic Acid and Anthraquinone-2,6-Disulfonate in Redox Systems ..................................................................................................... 154 Dehalorespiration................................................................................................................ 156 Alkane Sulfonates as Terminal Electron Acceptors........................................................... 156 Nitroalkanes as Electron Acceptor ..................................................................................... 156 References ........................................................................................................................... 157 Part 3 Reductases and Related Enzymes ....................................................................................... 162 Reduction of Nitroarenes .................................................................................................... 162 Nitrate Ester Reductase ...................................................................................................... 162 Enones (α, β-Unsaturated Ketones) and Related Reductases ............................................. 163 Aerobic Conditions .................................................................................................. 163 Anaerobic Conditions .............................................................................................. 163 Azo Reductase .................................................................................................................... 163

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Aldehyde Oxidoreductases ................................................................................................. 164 F420 -Dependent (Desazatetrahydrofolate) Reductases........................................................ 164 Other Carbocyclic Aromatic Compounds .......................................................................... 164 Reduction of Carbocyclic Rings .............................................................................. 164 Dehydroxylation ....................................................................................................... 165 Reduction of Azaarene Rings .................................................................................. 165 Metal Cations and Oxyanions ............................................................................................ 165 References ........................................................................................................................... 166 Part 4 Microbial Reactions to Chemical Stress ............................................................................. 168 Introduction ........................................................................................................................ 168 Hydrocarbon Tolerance............................................................................................ 168 Antibiotic Resistance ............................................................................................... 170 Resistance to Metals and Metalloids .................................................................................. 172 Reduction ................................................................................................................. 172 Methylation .............................................................................................................. 173 Efflux Systems ......................................................................................................... 175 Chlorophenol Tolerance ........................................................................................... 176 Resistance to High Acidity ...................................................................................... 176 References ........................................................................................................................... 177 Part 5 Enzymes Containing Manganese, Iron, Nickel, Copper, Molybdenum, Tungsten, and Vanadium ..................................................................................................... 181 Manganese .......................................................................................................................... 181 Iron .................................................................................................................................. 181 Nickel .................................................................................................................................. 182 Copper................................................................................................................................. 184 Molybdenum ....................................................................................................................... 185 Molybdopterin Oxidoreductases .............................................................................. 185 Tungsten .............................................................................................................................. 187 Vanadium ............................................................................................................................ 188 References ........................................................................................................................... 188 Chapter 4

Determinants and Interactions ................................................................................. 193

Single Substrates: Several Organisms ........................................................................................... 193 Cometabolism and Related Phenomena......................................................................................... 195 Induction of Catabolic Enzymes .................................................................................................... 195 Pre-Exposure to an Analog Substrate ................................................................................ 195 Enzyme Induction by Growth on Structurally Unrelated Compounds .............................. 197 Role of Readily Degraded Substrates ............................................................................................ 198 Physical Parameters .......................................................................................................................200 Temperature ........................................................................................................................200 Oxygen Concentration ........................................................................................................ 201 Anaerobic Bacteria ..................................................................................................202 Aerobic and Facultatively Anaerobic Organisms ....................................................202 Redox Potential ...................................................................................................................205 Association of Bacteria with Particulate Material: “Free” and “Bound” Substrates .............................................................................................................205 Biological Mechanisms for Association with Organic Components of Soil and Sediment ...........................................................................................................206 Aging ..................................................................................................................................208

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Substrate Concentration, Transport into Cells, and Toxicity ......................................................... 210 Utilization of Low Substrate Concentrations ..................................................................... 211 Existence of Threshold Concentrations .............................................................................. 212 Strategies Used by Cells for Substrates with Low or Negligible Water Solubility ............ 213 Transport Mechanisms ....................................................................................................... 214 Pre-Exposure: Pristine and Contaminated Environments ............................................................. 215 Rates of Metabolic Reaction .......................................................................................................... 218 Kinetic Aspects................................................................................................................... 218 Metabolic Aspects: Nutrients......................................................................................................... 219 Regulation and Toxic Metabolites ................................................................................................. 220 Regulation ........................................................................................................................... 220 Toxic or Inhibitory Metabolites .......................................................................................... 222 Catabolic Plasmids.........................................................................................................................224 References ...................................................................................................................................... 227

SECTION II Experimental Procedures Chapter 5

General Experimental Procedures ........................................................................... 245

Introduction.................................................................................................................................... 245 Abiotic Reactions ................................................................................................................ 245 Microbial Reactions ............................................................................................................246 Storage of Samples ............................................................................................................. 247 Determination of Ready Biodegradability ......................................................................... 247 The Inoculum ...........................................................................................................248 Concentration of the Substrate.................................................................................248 Endpoints .................................................................................................................248 Design of Experiments on Inherent Biodegradability ........................................................248 Metabolic Limitations..............................................................................................248 Application to Marine Systems................................................................................ 249 Isolation and Elective Enrichment ........................................................................... 249 General Procedures .................................................................................................. 250 Basal Media ............................................................................................................. 252 Mineral Media ......................................................................................................... 252 Trace Elements ......................................................................................................... 253 Control of pH ...........................................................................................................254 Vitamins...................................................................................................................254 Sterilization ..............................................................................................................254 Metal Concentration in Metabolism ........................................................................254 Redox Potential of Media ........................................................................................ 256 Organic Substrates .............................................................................................................. 256 Toxic Compounds .................................................................................................... 256 Volatile Compounds................................................................................................. 257 Solid Media ......................................................................................................................... 257 Growth at the Expense of Alternative Substrates ............................................................... 258 Techniques for Anaerobic Bacteria .................................................................................... 258 Design of Experiments on Biodegradation and Biotransformation ................................... 259 Pure Cultures and Stable Consortia.................................................................................... 259 Cell Growth at the Expense of the Xenobiotic ...................................................................260

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Stable Enrichment Cultures ................................................................................................260 Use of Dense Cell Suspensions........................................................................................... 261 Use of Immobilized Cells ................................................................................................... 261 Application of Continuous Culture Procedures.................................................................. 262 Simultaneous Presence of Two Substrates ......................................................................... 263 Use of Unenriched Cultures: Undefined Natural Consortia ...............................................264 Microcosm Experiments.....................................................................................................264 Experiments in Models of Natural Aquatic Systems .........................................................266 Evaluation of Degradation Using Metabolites.................................................................... 267 Experimental Problems: Water Solubility, Volatility, Sampling, and Association of the Substrate with Microbial Cells ........................................... 268 References ...................................................................................................................................... 269

Chapter 6

Elucidation of Metabolic Pathways .......................................................................... 277

Introduction.................................................................................................................................... 277 Part 1 Application of Natural and Synthetic Isotopes ................................................................... 277 Carbon (14C and 13C)........................................................................................................... 277 Sulfur (35S) and Chlorine (36Cl) .......................................................................................... 278 Hydrogen (2H) and Oxygen (18O) ........................................................................................ 278 Other Isotopes .....................................................................................................................280 Isotope Effects and Stable Isotope Fractionation ...............................................................280 Experimental Procedures......................................................................................... 282 References ........................................................................................................................... 282 Part 2 Nondestructive Methods for Structure Determination .......................................................284 Application of NMR ...........................................................................................................284 Hydrogen 1H ............................................................................................................ 285 Carbon 13C ............................................................................................................... 285 Nitrogen 15N ............................................................................................................. 286 Oxygen 17O ............................................................................................................... 287 Fluorine 19F .............................................................................................................. 287 Phosphorus 31P ......................................................................................................... 288 Silicon 29Si ............................................................................................................... 288 Application of EPR ............................................................................................................. 288 Elucidation of the Unexpected Complexity of an Apparently Straightforward Metabolic Pathway ..................................................................................... 289 Hydroxyl Radicals and Their Role in the Toxicity of H2O2 to Bacteria ......................................................................................... 289 Elucidation of the Mechanism of Pyruvate Formate Lyase........................................................................................................... 289 The Function of Humic Acids in Reactions Catalyzed by Geobacter metallireducens .......................................................................................... 289 The Mechanism of Anaerobic Activation of Toluene .............................................. 289 Dioxygenation with Elimination of Halide .............................................................. 289 The Anaerobic Degradation of Benzoate ................................................................290 Manganese-Containing Enzymes ............................................................................ 290 Turnover of Naphthalene Dioxygenase ....................................................................290 Enzyme Structure ...............................................................................................................290 References ........................................................................................................................... 291

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SECTION III Pathways and Mechanisms of Degradation and Transformation Chapter 7

Aliphatic Compounds............................................................................................... 297

Part 1 Alkanes and Related Compounds ....................................................................................... 297 Alkanes ............................................................................................................................... 297 Aerobic Conditions .................................................................................................. 297 Anaerobic Conditions ..............................................................................................304 Alkenes ............................................................................................................................... 305 Aerobic Conditions .................................................................................................. 305 Anaerobic Conditions ..............................................................................................308 Alkynes ...............................................................................................................................308 Alkanols ..............................................................................................................................309 Aerobic Conditions ..................................................................................................309 Anaerobic Conditions .............................................................................................. 310 Alkanones ........................................................................................................................... 310 Alkylamines and Amino Acids .......................................................................................... 311 Aerobic Conditions .................................................................................................. 311 Anaerobic Conditions .............................................................................................. 316 Alkanoic Acids ................................................................................................................... 317 Aerobic Conditions .................................................................................................. 317 Anaerobic Conditions .............................................................................................. 318 Anaerobic Conditions .............................................................................................. 318 Amides and Related Compounds ....................................................................................... 321 Amides .................................................................................................................... 321 Nitriles ................................................................................................................................ 322 Isonitriles ............................................................................................................................ 322 Sulfonylureas and Thiocarbamates .................................................................................... 322 Carbon Monoxide ............................................................................................................... 323 Cyanide ............................................................................................................................... 323 Thiocyanate ........................................................................................................................ 324 References ........................................................................................................................... 324 Part 2 Cycloalkanes ....................................................................................................................... 336 Monoterpenes ..................................................................................................................... 339 Steroids ............................................................................................................................... 341 Hydroxylation........................................................................................................... 341 Other Aerobic Transformations and Degradation ................................................... 341 Anaerobic Transformation ....................................................................................... 343 Diterpenoids........................................................................................................................344 Polycyclic Substrates ..........................................................................................................344 References ...........................................................................................................................346 Part 3 Alkanes, Cycloalkanes and Related Compounds with Chlorine, Bromine, or Iodine Substituents ......................................................................................................................... 349 Chlorinated, Brominated, and Iodinated Alkanes, Alkenes, and Alkanoates ................... 349 Elimination Reactions.............................................................................................. 350 Corrinoid Pathways .................................................................................................. 355 Nucleophilic Substitution: Hydrolytic Reactions of Halogenated Alkanes and Alkanoates ....................................................... 358

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Glutathione-Mediated and Other Reactions Involving Nucleophilic Sulfur .................................................................................... 363 Monooxygenation..................................................................................................... 363 Reductive Reactions: Dehalogenation Including Dehalorespiration ....................... 366 References ........................................................................................................................... 370 Part 4 Fluorinated Aliphatic Compounds ...................................................................................... 378 Alkanes and Alkenes .......................................................................................................... 378 Fluorohydrocarbons ................................................................................................. 378 Chlorofluorocarbons and Hydrochlorofluorocarbons .............................................. 379 Carboxylic Acids...................................................................................................... 380 Perfluoroalkyl Carboxylates and Sulfonates ............................................................ 382 References ........................................................................................................................... 382

Chapter 8

Carbocyclic Aromatic Compounds without Halogen Substituents .......................... 385

Part 1 Monocyclic Aromatic Hydrocarbons .................................................................................. 385 Introduction ........................................................................................................................ 385 Monocyclic Arenes ............................................................................................................. 385 Aerobic Conditions .................................................................................................. 385 Anaerobic Conditions .............................................................................................. 389 Synthetic Applications ............................................................................................. 392 References ........................................................................................................................... 394 Part 2 Polycyclic Aromatic Hydrocarbons (PAHs)........................................................................ 398 Introduction ........................................................................................................................ 398 Aerobic Reactions Carried Out by Bacteria ............................................................ 399 PAHs with Three or More Rings .............................................................................402 Anaerobic Reactions Carried Out by Bacteria ........................................................408 Fungal Transformations ...........................................................................................409 White-Rot Fungi ...................................................................................................... 413 References ........................................................................................................................... 417 Part 3 Aromatic Carboxylates, Carboxaldehydes, and Related Compounds ................................ 424 Introduction ........................................................................................................................ 424 Benzoates ............................................................................................................................ 424 Aerobic Conditions .................................................................................................. 424 Hydroxybenzoates and Related Compounds ...................................................................... 425 Mechanisms for Fission of Oxygenated Rings ........................................................ 428 Alternative Pathways for the Degradation of Benzoates and Related Compounds............................................................................. 432 Aerobic Reduction of Arene Carboxylates ......................................................................... 433 Arenes with an Oxygenated C2 or C3 Side Chain ............................................................... 433 Anaerobic Metabolism ............................................................................................. 435 Aldehydes ........................................................................................................................... 439 References ........................................................................................................................... 439 Part 4 Nonhalogenated Phenols and Anilines ...............................................................................446 Phenols ................................................................................................................................446 Aerobic Degradation ................................................................................................446 Anaerobic Degradation ............................................................................................449 Anilines.................................................................................................................... 452 References ........................................................................................................................... 452

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Chapter 9

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Substituted Carbocyclic Aromatic Compounds ....................................................... 455

Part 1 Halogenated Arenes and Carboxylates with Chlorine, Bromine, or Iodine Substituents.......................................................................................................... 455 Introduction ........................................................................................................................ 455 Halogenated Arene Hydrocarbons ..................................................................................... 455 Aerobic Conditions .................................................................................................. 455 Monocyclic Chlorinated Arenes ......................................................................................... 456 Anaerobic Conditions .............................................................................................. 458 Polychlorinated Biphenyls .................................................................................................. 458 Aerobic Degradation ................................................................................................ 458 Degradation Is Initiated by Dioxygenation .............................................................. 459 Ring Fission by 2,3-Dihydroxybiphenyl Dioxygenase ............................................ 461 Metabolites............................................................................................................... 463 Fungal Dehalogenation .......................................................................................................465 Reductive Dehalogenation ..................................................................................................465 The Role of Temperature .........................................................................................466 Polyhalogenated Dioxins and Diphenyl Ethers........................................................468 Halogenated Benzoates.......................................................................................................468 Dioxygenation ..........................................................................................................469 Hydrolytic Reactions................................................................................................ 472 Mechanisms for the Ring Fission of Substituted Catechols ............................................... 472 Reductive Loss of Halogen ...................................................................................... 474 Halogenated Phenylacetates ............................................................................................... 475 Fungal Reactions...................................................................................................... 476 References ........................................................................................................................... 476 Part 2 Halogenated (Chlorine, Bromine, and Iodine) Phenols and Anilines ........................................................................................................................ 482 Phenols ................................................................................................................................ 482 Aerobic Conditions .................................................................................................. 482 Fungi and Yeasts ...................................................................................................... 486 Anaerobic Conditions .............................................................................................. 487 Anilines .............................................................................................................................. 489 Aerobic Conditions .................................................................................................. 489 Anaerobic Conditions .............................................................................................. 490 References ........................................................................................................................... 491 Part 3 Fluorinated Hydrocarbons, Carboxylates, Phenols, and Anilines ...................................... 494 Fluorinated Aromatic Hydrocarbons.................................................................................. 494 Aerobic Conditions .................................................................................................. 494 Metabolism by Yeasts and Fungi ............................................................................. 495 Anaerobic Denitrifying Conditions ......................................................................... 495 Fluorobenzoates .................................................................................................................. 496 Aerobic Conditions .................................................................................................. 496 Difluorobenzoates .................................................................................................... 498 Degradation under Denitrifying Conditions ............................................................ 499 Fluorinated Phenols ............................................................................................................500 Aerobic Conditions ..................................................................................................500 Anaerobic Conditions .............................................................................................. 501 Aromatic Trifluoromethyl Compounds ..............................................................................502 References ........................................................................................................................... 503

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Part 4 Arene Sulfonates .................................................................................................................506 References ...........................................................................................................................508 Part 5 Aromatic Compounds with Nitro Substituents ...................................................................508 Nitroarenes..........................................................................................................................508 Reduction of Nitro Groups.......................................................................................509 Reduction of the Ring .............................................................................................. 512 Dioxygenation .......................................................................................................... 512 Side-Chain Oxidation .............................................................................................. 513 Peroxidase Oxidation ............................................................................................... 513 Nitrobenzoates .................................................................................................................... 514 Nitrophenols........................................................................................................................ 514 References ........................................................................................................................... 517 Part 6 Azoarenes ............................................................................................................................ 520 References ........................................................................................................................... 521 Chapter 10

Heterocyclic Aromatic Compounds ....................................................................... 523

Part 1 Azaarenes ............................................................................................................................ 523 Five-Membered Monocyclic Aza, Oxa, and Thiaarenes .................................................... 523 Aerobic Conditions .................................................................................................. 523 Indole and Carbazole .......................................................................................................... 524 Indole and 3-Alkylindoles ....................................................................................... 524 Carbazole ................................................................................................................. 527 Pyridine .............................................................................................................................. 527 Aerobic Conditions .................................................................................................. 527 Anaerobic Conditions .............................................................................................. 534 Quinoline and Isoquinoline ................................................................................................ 536 Bacterial Metabolism ............................................................................................... 536 Hydroxylation........................................................................................................... 537 Dioxygenation .......................................................................................................... 537 Pyrimidine and Related Compounds Excluding Purines ...................................................540 Aerobic Conditions ..................................................................................................540 Anaerobic Conditions .............................................................................................. 541 Purines ................................................................................................................................ 542 Aerobic Degradation ................................................................................................ 542 Anaerobic Degradation ............................................................................................ 542 Ring-Fission Reactions ............................................................................................544 Triazines ............................................................................................................................. 545 1,3,5-Triazines.......................................................................................................... 545 1,2,4-Triazines..........................................................................................................546 References ...........................................................................................................................546 Part 2 Oxaarenes ............................................................................................................................ 553 Aerobic Conditions ............................................................................................................. 553 Monocyclic Oxaarenes............................................................................................. 553 Polycyclic Oxaarenes ............................................................................................... 553 References ........................................................................................................................... 562 Part 3 Thiaarenes: Benzothiophenes, Dibenzothiophenes, and Benzothiazole ............................ 565 Benzothiazole ..................................................................................................................... 567 References ........................................................................................................................... 568

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Chapter 11 Miscellaneous Compounds ...................................................................................... 569 Part 1 Carboxylate, Sulfate, Phosphate, and Nitrate Esters ........................................................... 569 Carboxylates ....................................................................................................................... 569 Sulfates ............................................................................................................................... 569 Phosphates .......................................................................................................................... 570 Nitrates................................................................................................................................ 571 References ........................................................................................................................... 572 Part 2 Ethers and Sulfides .............................................................................................................. 573 Aliphatic and Benzylic Ethers ............................................................................................ 573 Aryl Ethers.......................................................................................................................... 576 Diaryl Ethers ............................................................................................................ 576 Aryl-Alkyl Ethers .................................................................................................... 576 Sulfides, Disulfides, and Related Compounds .................................................................... 578 References ........................................................................................................................... 581 Part 3 Aliphatic Nitramines and Nitroalkanes .............................................................................. 585 Nitramines .......................................................................................................................... 585 Nitroalkanes........................................................................................................................ 585 References ........................................................................................................................... 586 Part 4 Aliphatic Phosphonates and Sulfonates .............................................................................. 588 Introduction ........................................................................................................................ 588 Phosphonates ...................................................................................................................... 588 Sulfonates ........................................................................................................................... 589 Boronates ............................................................................................................................ 591 References ........................................................................................................................... 591 Part 5 Degradation of Organic Compounds of Metals and Metalloids ......................................... 592 Tin .................................................................................................................................. 593 Lead .................................................................................................................................. 593 Mercury .............................................................................................................................. 594 Arsenic ................................................................................................................................ 594 References ........................................................................................................................... 594

SECTION IV

Bioremediation

Chapter 12 Bioremediation: Overview ....................................................................................... 599 Introduction.................................................................................................................................... 599 Strategies............................................................................................................................. 599 Contaminants ......................................................................................................................600 Sites ..................................................................................................................................602 Chemical Procedures ..........................................................................................................602 Phytoremediation ................................................................................................................602 Primary Role of Plants.............................................................................................603 Secondary Role of Plants .........................................................................................605 Degradative Enzymes ..............................................................................................606 The Role of Plant Exudates in Degradation ............................................................607 Other Functions of Bacteria in the Rhizosphere .....................................................607 The Role of Mycorrhizal Fungi ...............................................................................609

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Microbiological Aspects ................................................................................................................ 610 Regulation of Pathways ...................................................................................................... 610 Alternative Electron Acceptors .......................................................................................... 611 Aging .................................................................................................................................. 611 Biofilms............................................................................................................................... 611 Metabolites and Enzymes ................................................................................................... 611 Utilization of Nitrogen, Sulfur, and Phosphorus ................................................................ 612 Substrate Concentration ...................................................................................................... 612 Temperature ........................................................................................................................ 613 References ...................................................................................................................................... 613

Chapter 13 Evaluation of Bioremediation ................................................................................... 621 Introduction.................................................................................................................................... 621 Analysis of Populations ................................................................................................................. 622 Analysis of Degradative Populations.................................................................................. 622 Application to the Degradation of Specific Contaminants................................................. 623 Hydrocarbons ........................................................................................................... 623 Trichloroethene ........................................................................................................ 624 Phenol .................................................................................................................... 624 Chlorophenol............................................................................................................ 625 Chlorobenzoate ........................................................................................................ 625 Phenylurea Herbicides ............................................................................................. 625 Dehalogenation of Chloroalkanoates....................................................................... 626 Dechlorination of PCBs ........................................................................................... 626 Application to Specific Groups of Organisms .................................................................... 626 Nondirected Examination of Natural Populations ............................................................. 627 Application of Stable Isotopes ....................................................................................................... 627 Stable Isotope Probes.......................................................................................................... 627 Application of Stable Isotope Enrichment.......................................................................... 628 Application of δ13C .................................................................................................. 629 References ...................................................................................................................................... 633

Chapter 14 Applications of Bioremediation ............................................................................... 639 Introduction.................................................................................................................................... 639 Part 1 Petroleum Hydrocarbons: Refinery Waste and Stranded Oil.............................................. 639 Terrestrial Habitats .............................................................................................................640 Alkanes ....................................................................................................................640 Naphthenates ............................................................................................................ 641 Marine Habitats .................................................................................................................. 641 The Baffin Island Oil Spill Project .......................................................................... 641 Conclusion .......................................................................................................................... 642 References ........................................................................................................................... 642 Part 2 Coal Distillation Products ................................................................................................... 643 Introduction ........................................................................................................................ 643 On-Site or Off-Site Treatment ............................................................................................644 The Physical State of the Contaminant ..............................................................................646

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Environmental Parameters ................................................................................................. 647 Determinants ...................................................................................................................... 647 The Range of Substrates .......................................................................................... 647 The Range of Degradative Bacteria .........................................................................648 Bioavailability and the Use of Surfactants ..............................................................649 Supplementation with Fungi ............................................................................................... 650 Application of Higher Plants .............................................................................................. 652 Anoxic or Anaerobic Environments ................................................................................... 652 Hydrocarbons ........................................................................................................... 652 Phenols .................................................................................................................... 652 Heteroarenes ............................................................................................................ 653 Abiotic Transformations ..................................................................................................... 653 Conclusions ......................................................................................................................... 654 References ........................................................................................................................... 654 Part 3 Wood Preservation Sites: Chlorophenolic Compounds ...................................................... 658 Aerobic Bacteria ................................................................................................................. 659 Degradation by Fungi ......................................................................................................... 659 Anaerobic Dechlorination ..................................................................................................660 Conclusion ..........................................................................................................................660 References ...........................................................................................................................660 Part 4 Chemical Waste................................................................................................................... 661 Chlorinated Alicyclic Hydrocarbons .................................................................................. 662 Chlorinated and Brominated Aromatic Hydrocarbons ...................................................... 662 Chlorobenzenes .................................................................................................................. 662 DDT .................................................................................................................................. 663 Polychlorinated Biphenyls .................................................................................................. 663 Aerobic Conditions ..................................................................................................664 Anaerobic Conditions ..............................................................................................665 Polybrominated Biphenyls and Diphenyl Ethers ................................................................666 Chlorinated Dibenzodioxins...............................................................................................666 Organofluoro Compounds .................................................................................................. 667 Conclusion ..........................................................................................................................668 References ...........................................................................................................................668 Part 5 Agrochemicals .................................................................................................................... 671 Phenoxyalkanoic Acids....................................................................................................... 672 Chlorinated Anilines .......................................................................................................... 672 Triazines ............................................................................................................................. 673 1,3,5-Triazines.......................................................................................................... 673 1,2,4-Triazines.......................................................................................................... 673 2,4-Dintro-6-sec-Butylphenol (Dinoseb)................................................................. 673 References ........................................................................................................................... 674 Part 6 Military Waste..................................................................................................................... 675 Explosives ........................................................................................................................... 675 Nitroarenes.......................................................................................................................... 675 Nitrodiphenylamines .......................................................................................................... 676 Nitrate Esters ...................................................................................................................... 676 Nitramines .......................................................................................................................... 676 Conclusions ......................................................................................................................... 677 Chemical Warfare Agents .................................................................................................. 677 References ........................................................................................................................... 677

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Contents

Part 7 Groundwater Contamination ............................................................................................... 679 Benzene/Toluene/Ethylbenzene/Xylenes............................................................................ 679 Aerobic Conditions .................................................................................................. 679 Anoxic or Anaerobic Conditions .............................................................................680 Halogenated Alkanes and Alkenes..................................................................................... 682 Tetrachloromethane ................................................................................................. 682 Chloroethenes .......................................................................................................... 682 Aerobic Conditions .................................................................................................. 682 Application of Indigenous Bacteria ......................................................................... 683 Application of Exogenous Bacteria .........................................................................684 Anoxic and Anaerobic Conditions ...........................................................................684 Application of Higher Plants ................................................................................... 685 Conclusion .......................................................................................................................... 686 Methyl tert-Butyl Ether....................................................................................................... 686 References ........................................................................................................................... 687 Finale: A Hierarchical Strategy .....................................................................................................690 I. Laboratory Experiments..................................................................................................690 II. Microcosm Experiments ................................................................................................ 691 III. Large-Scale Operations ................................................................................................ 691 Concluding Comments................................................................................................................... 692 Index .............................................................................................................................................. 693

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Preface This volume is an updated and expanded version of a previous edition Organic Chemicals: An Environmental Perspective. This one deals, however, with only degradation and transformation in their widest senses, and the sections in the earlier volumes on analysis, distribution, and ecotoxicology have been omitted since these lie beyond my current competence. In addition, there are specialized volumes that already cover these topics. I have been extremely fortunate in having as coauthor Ann-Sofie Allard, who has been a research collaborator for many years. She has been both guide and illustrator, and without her constant help and encouragement, this volume would neither have begun nor would it have come to fruition. Degradation and transformation occupy a central position in assessing the environmental impact of organic contaminants. We have avoided the term “pollutant” on account of its negative connotation. Although phase partition is not discussed, this is an important factor in determining biodegradability since a contaminant will seldom remain in the phase to which it is initially discharged. This must be considered as the following examples illustrate: 1. Compounds deposited as solids (including agrochemicals and contaminants) may reach groundwater and watercourses as a result of partition and leaching. 2. Substances with even marginal volatility will enter the atmosphere, and after transformation may then reenter the aquatic and terrestrial environments through precipitation. 3. Aquatic biota may bring about transformation to metabolites that are then disseminated to a considerable distance from their source. 4. Particularly polar contaminants may associate with polymeric humic components of soil, water, and sediment. Their biodegradation then depends on the degree to which these processes are reversible and the contaminants become accessible to microorganisms (bioavailable). This is especially significant after weathering (aging), even for nonpolar compounds. Contaminants seldom consist of single substances. The pathways used for biodegradation of some components may be incompatible with those for others that are present, or the pathway for a single compound. Examples of this are given, though in less detail than they merit. In this volume, emphasis is placed on the pathways by which degradation or transformation has taken place, and the approach is essentially chemical and mechanistic. There are several reasons for this. 1. Biochemical reactions parallel those in organic chemistry and, for both of them, a mechanistic approach has proved valuable. In addition, most of the principles that have emerged apply equally to the aquatic, the atmospheric, and the terrestrial environments. 2. Metabolites may be produced by biochemical transformation of the substrate rather than by degradation, or may result from partial abiotic reactions. These products may be (a) terminal and persistent or (b) toxic to other components of an ecosystem—including the microorganisms that produce them. Both of these represent important considerations that are illustrated by examples in this book. 3. Pathways can provide a guide to the probable reactions that other contaminants may undergo including those that bear only a partial—or even more remote structural resemblance. This is especially important in view of the widening spectrum of contaminants that is the result of the impressive ingenuity of the organic chemist, and particularly for the complexity and novelty of pharmaceutical and agrochemical products. xix

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xx

Preface

A word of caution is appropriate. The reactions that are used for illustration are very seldom specific for a single taxon—or even close relatives, and it is not generally possible to establish the range of organisms that will be able to carry out the reaction. The contents of the chapters are interdependent. In summary they deal with the following: Chapters 1 through 4 provide a broad perspective on abiotic and biotic reactions, including the significance of a range of environmental determinants. Chapters 5 and 6 attempt to provide a brief introduction to experimental procedures with emphasis on procedures for establishing the structure of metabolites using isotopes and physical methods. Chapters 7 through 11 outline details of biochemical reactions involved in the biodegradation of the major groups of aliphatic, carbocyclic aromatic, and heterocyclic compounds. Although emphasis is placed on the pathways, rather general accounts of the enzymes involved and the genetics are provided where they are available. Chapters 12 through 14 deal with bioremediation that has attracted increasing concern with the realization of the hazard presented by the disposal of unwanted chemicals, or byproducts from their manufacture. These chapters should be viewed within the wider context of metabolic details that have been presented in Chapters 7 through 11. There are a number of inevitable omissions and limitations in the material that is covered. 1. Emphasis is placed exclusively on xenobiotics, although some of them are also naturally occurring metabolites. There are substantial groups of compounds that are not discussed. These include the following: a. Naturally occurring polymers—cellulose, chitin, lignin, and polyisoprenoids; synthetic polyamides, polyurethanes, and polysiloxanes. b. Natural products including polypyrroles such as hemin and chlorophyll; plant and microbial metabolites. c. Agrochemicals including veterinary chemotherapeutic agents that have come into use in large-scale animal husbandry. d. Pharmaceuticals including hormone disrupters about which there has arisen serious concern. e. Halogenated metabolites produced by marine biota, even though some such as halomethanes, polybrominated phenols, and polybrominated diphenyl ethers are also xenobiotics. It is expected that the pharmaceutical industry that has hitherto been dominated by synthetic products—except for antibiotics—will come to rely increasingly on metabolites produced by marine biota including microorganisms. 2. The number of references that are cited is necessarily restrictive and, although numerous, they represent merely an eclectic selection from a vast literature. Relevant references have no doubt been omitted, but the writers can assure the authors of these that there is no malice in the selection. They invariably refer to the primary literature that has been subjected to the scrutiny of peer review. It is therefore assured that even when the interpretations of the authors should prove faulty—and this is inevitable—a solid and reproducible basis of fact is available to the critical reader. Some older work has been cited when this has led to lasting concepts, though other early work may be difficult to evaluate by the standards of today; and no doubt work at the cutting edge of current research will rightly require modification and extension in the future. 3. Although some of the examples used for illustration in Chapter 14 impinge on biological wastewater treatment technology, a systematic account of these lies beyond the scope of this volume and the competence of the authors. It is worth noting, however, that anaerobic reactors that are developed to treat wastewater with the object of producing methane may

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Preface

xxi

not necessarily degrade recalcitrant xenobiotics for which inoculation with specific microorganisms may be necessary. 4. No discussion of models will be found in this volume, either for the analysis of degradation kinetics or for the prediction of biodegradability. For these, the interested reader should consult monographs by experts. The nomenclature of bacteria has presented serious problems principally for two reasons: (a) the number of new taxa that are being described appears to be increasing exponentially and (b) it is seldom possible to determine the correct taxonomic assignment for important organisms that have been used in the historical literature. Although no consistent attempt has been made to provide the current assignment for all taxa, we have tried to use those that occur most frequently.

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Acknowledgments It is a pleasure to thank Östen Ekengren, director of Environmental Technology and Toxicology for his generosity in extending access to library facilities. We thank Springer Verlag for kind permission to use a number of figures that were used in chapters that we have contributed to Volume 3J (1998), Volume 3N (2002), and Volume 3R (2003) of The Handbook of Environmental Chemistry of which I was volume editor.

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Authors Alasdair H. Neilson was principal scientist until retirement from IVL Swedish Environmental Research Institute in Stockholm. He studied chemistry at the University of Glasgow and obtained his PhD in organic chemistry from Alexander Todd’s laboratory at Cambridge. He carried out research at Cambridge in organic chemistry, and in theoretical chemistry with Charles Coulson at Oxford. He held academic positions in the universities of Glasgow and Sussex, and obtained industrial experience in the pharmaceutical industry. He consolidated his experience by turning to research in microbiology during a prolonged stay with Roger Stanier and Mike Doudoroff in Berkeley. His interests have ranged widely and included studies on nitrogen fixation, carbon and nitrogen metabolism in algae, and various aspects of environmental science including biodegradation and biotransformation, chemical and microbiological reactions in contaminated sediments, and ecotoxicology. With his group of collaborators, these studies have resulted in publications in Applied and Environmental Microbiology, Journal of Chromatography, Environmental Science & Technology, and Ecotoxicology & Environmental Safety, and in chapters contributed to several volumes of The Handbook of Environmental Chemistry. He is a member of the American Chemical Society, the American Society for Microbiology, the American Society of Crystallography, and the AAAS. Ann-Sofie Allard was trained as a chemical microbiologist and is currently a senior microbiologist at IVL Swedish Environmental Research Institute in Stockholm. She has carried out research in a wide range of environmental issues including water quality, processes for the removal of hormone disrupters, biodegradation and biotransformation of organic contaminants in aquatic and terrestrial systems, and ecotoxicology. She has implemented studies on the uptake and metabolism of organic contaminants and metals in higher plants in the context of bioremediation. Her studies have been published in Applied and Environmental Microbiology, International Biodeterioration and Biodegradation, Environmental Chemistry and Ecotoxicology, and Journal of Environmental Science and Health, and in chapters contributed to several volumes of The Handbook of Environmental Chemistry. She is a member of the American Chemical Society.

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Section I Degradation and Transformation Processes

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1

Abiotic Reactions

INTRODUCTION Virtually any of the plethora of reactions in organic chemistry may be exploited for the abiotic degradation of xenobiotics. These include nucleophilic displacement, oxidation, reduction, thermal reactions, and halogenation. Hydrolytic reactions may convert compounds such as esters, amides or nitriles into the corresponding carboxylic acids, or ureas and carbamides into the amines. These abiotic reactions may therefore be the first step in the degradation of such compounds. The transformation products may, however, be resistant to further chemical transformation so that their ultimate fate is dependent upon subsequent microbial reactions. For example, for some urea herbicides, the limiting factor is the rate of microbial degradation of the chlorinated anilines that are the initial products of hydrolysis. The role of abiotic reactions should therefore always be taken into consideration, and should be carefully evaluated in laboratory experiments on biodegradation and biotransformation. The results of experiments directed to microbial degradation are probably discarded if they show substantial interference from abiotic reactions. A good illustration of the complementary roles of abiotic and biotic processes is offered by the degradation of tributyltin compounds. Earlier experiments (Seligman et al. 1986) had demonstrated the degradation of tributyltin to dibutyltin primarily by microbial processes. It was subsequently shown, however, that an important abiotic reaction mediated by fine-grained sediments resulted in the formation of monobutyltin and inorganic tin also (Stang et al. 1992). It was therefore concluded that both processes were important in determining the fate of tributyltin in the marine environment. A study of the carbamate biocides, carbaryl, and propham illustrates the care that should be exercised in determining the relative importance of chemical hydrolysis, photolysis, and bacterial degradation (Figure 1.1) (Wolfe et al. 1978). For carbaryl, the half-life for hydrolysis increased from 0.15 d at pH 9 to 1500 d at pH 5, while that for photolysis was 6.6 d: biodegradation was too slow to be significant. In contrast, the half-lives of propham for hydrolysis and photolysis were >104 and 121 d—so greatly exceeding the half-life of 2.9 d for biodegradation that abiotic processes would be considered to be of subordinate significance. Close attention to structural features of xenobiotics is therefore clearly imperative before making generalizations on the relative significance of alternative degradative pathways.

PHOTOCHEMICAL REACTIONS IN AQUEOUS AND TERRESTRIAL ENVIRONMENTS Photochemical reactions are important in atmospheric reactions, in terrestrial areas of high solar irradiation such as the surface of soils, and in aquatic systems containing ultraviolet (UV)absorbing humic and fulvic acids (Zepp et al. 1981a,b). They may be relevant especially for otherwise recalcitrant compounds. It has also been shown (Zepp and Schlotzhauer 1983) that although the presence of algae may enhance photometabolism, this is subservient to direct photolysis at the cell densities likely to be encountered in rivers and lakes. It should be noted that different products may be produced in natural river water and in buffered medium. For example, photolysis of triclopyr (3,5,6-trichloro-2-pyridyloxyacetic acid) in sterile medium at pH 7 resulted in hydrolytic replacement of one chlorine atom, whereas in river water the ring was degraded to form oxamic acid as the principal product (Woodburn et al. 1993). Particular attention has therefore been understandably directed to the photolytic degradation of biocides—including agrochemicals—that are applied to terrestrial systems and enter the aquatic system through leaching. There has been increased interest 3

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4

Environmental Degradation and Transformation of Organic Chemicals O−CO−NHCH3

NH−CO−O−CH(CH3)2

(A)

FIGURE 1.1

(B)

Carbaryl (A) and propham (B). OH O

O O

O

O

O

O HO 2 C

O O O

FIGURE 1.2

Photochemical transformation of santonin.

in their phototoxicity toward a range of biota (references in Monson et al. 1999), and this may be attributed to some of the reactions and transformations that are discussed later in this chapter. It should be emphasized that photochemical reactions may produce molecules structurally more complex and less susceptible to degradation than their precursors, even though the deep-seated rearrangements induced in complex compounds such as the terpene santonin during UV irradiation (Figure 1.2) are not likely to be encountered in environmental situations.

THE DIVERSITY OF PHOTOCHEMICAL TRANSFORMATIONS In broad terms, the following types of reactions are mediated by the homolytic fission products of water (formally, hydrogen, and hydroxyl radicals), and by molecular oxygen including its excited states—hydrolysis, elimination, oxidation, reduction, and cyclization.

THE ROLE OF HYDROXYL RADICALS The hydroxyl radical plays two essentially different roles: (a) as a reactant mediating the transformations of xenobiotics and (b) as a toxicant that damages DNA. They are important in a number of environments: (1) in aquatic systems under irradiation, (2) in the troposphere, which is discussed later, and (3) in biological systems in the context of superoxide dismutase and the role of iron. Hydroxyl radicals in aqueous media can be generated by several mechanisms: a. Photolysis of nitrite and nitrate (Brezonik and Fulkerson-Brekken 1998) b. Fenton reaction with H2O2 and Fe2+ in the absence or presence (Fukushima and Tatsumi 2001) of light c. Photolysis of fulvic acids under anaerobic conditions (Vaughan and Blough 1998) d. Reaction of Fe(III) or Cu(II) complexes of humic acids with hydrogen peroxide (Paciolla et al. 1999).

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Abiotic Reactions

5

ILLUSTRATIVE EXAMPLES OF PHOTOCHEMICAL TRANSFORMATIONS IN AQUEOUS SOLUTIONS 1. Atrazine is successively transformed to 2,4,6-trihydroxy-1,3,5-triazine (Pelizzetti et al. 1990) by dealkylation of the alkylamine side chains and hydrolytic displacement of the ring chlorine and amino groups (Figure 1.3). A comparison has been made between direct photolysis and nitrate-mediated hydroxyl radical reactions (Torrents et al. 1997): the rates of the latter were much greater under the conditions of this experiment, and the major difference in the products was the absence of ring hydroxylation with loss of chloride. 2. Pentachlorophenol produces a wide variety of transformation products, including chloranilic acid (2,5-dichloro-3,6-dihydroxybenzo-1,4-quinone) by hydrolysis and oxidation, a dichlorocyclopentanedione by ring contraction, and dichloromaleic acid by cleavage of the aromatic ring (Figure 1.4) (Wong and Crosby 1981). 3. The main products of photolysis of 3-trifluoromethyl-4-nitrophenol are 2,5 dihydroxybenzoate produced by hydrolytic loss of the nitro group and oxidation of the trifluoromethyl group, together with a compound identified as a condensation product of the original compound and the dihydroxybenzoate (Figure 1.5) (Carey and Cox 1981). Cl N C 2 H 5 NH

FIGURE 1.3

Cl

Cl N

N NHC 3 H 7 i

N

H2N

N N

N NH 2

N N

HO

OH

Photochemical transformation of atrazine. OH Cl

OH

Cl

Cl

O

Cl OH Cl

OH

Cl

O Cl

Cl

Cl

HO

Cl

O Cl

Cl

Cl

Cl

Cl

Cl

CO2H

Cl

CO2H

O OH

FIGURE 1.4

Cl

Cl

Cl

Cl

Photochemical transformation of pentachlorophenol.

CF3

CF3 NO2

FIGURE 1.5

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O

OH

OH

OH

OH

CO2H

CF3 O

OH

Photochemical transformation of 3-trifluoromethyl-4-nitrophenol.

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6

Environmental Degradation and Transformation of Organic Chemicals

4. A potential insecticide that is a derivative of tetrahydro-1,3-thiazine undergoes a number of reactions resulting in some 43 products of which the dimeric azo compound is the principal one in aqueous solutions (Figure 1.6) (Kleier et al. 1985). 5. The herbicide trifluralin undergoes a photochemical reaction in which the n-propyl side chain of the amine reacts with the vicinal nitro group to form the benzopyrazine (Figure 1.7) (Soderquist et al. 1975). 6. Heptachlor and cis-chlordane, both of which are chiral, produce caged or half-caged structures (Figure 1.8) on irradiation, and these products have been identified in biota from the Baltic, the Arctic, and the Antarctic (Buser and Müller 1993). 7. Methylcyclopentadienyl manganese tricarbonyl that has been suggested as a fuel additive is decomposed primarily by photolysis in aqueous medium. This resulted in the formation of methylcyclopentadiene that may plausibly be presumed to polymerize, and a manganese carbonyl that decomposed to Mn3O4 (Garrison et al. 1995). 8. Stilbenes that are used as fluorescent whitening agents are photolytically degraded by reactions involving cis–trans isomerization followed by hydration of the double bond, or oxidative fission of the double bond to yield aldehydes (Kramer et al. 1996). 9. The photolysis of chloroalkanes and chloroalkenes has received considerable attention and results in the formation of phosgene as one of the final products. The photodegradation of 1,1,1-trichloroethane involves hydrogen abstraction and oxidation to trichloroacetaldehyde that is degraded by a complex series of reactions to phosgene (Platz et al. 1995; Nelson et al. 1990). Tetrachloroethene is degraded by reaction with chlorine radicals and oxidation to pentachloropropanol radical that also forms phosgene (Franklin 1994). Attention is drawn to these reactions in the context of the atmospheric dissemination of xenobiotics. 10. Although ethylenediaminetetraacetic acid (EDTA) is biodegradable under specific laboratory conditions (Belly et al. 1975; Lauff et al. 1990; Nörtemann 1992; Witschel et al. 1997), NO2

N

N

N

S

S

N C

+ S

FIGURE 1.6

N

Photochemical transformation of tetrahydro-1,3-thiazine. Pr n

Pr n

Pr n

N

Et

H

NO 2

N N+

O 2N

CF 3

Photochemical transformation of trifluralin.

Cl

Cl

Cl Cl

Cl

Cl

Cl

Cl Cl

Cl Cl

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N

O 2N

O: −

CF 3

CF 3

FIGURE 1.8

Et

N

O 2N

FIGURE 1.7

N

N

S

Cl

Cl Cl

Photochemical transformation of chlordane.

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Abiotic Reactions

7 O

N Et

O

H 2N

N

O

HO2 C

11.

12.

13.

14.

15.

16.

N

N

O CO 2H

HO 2 C

FIGURE 1.9

CO 2 H

F

CO 2 H

F

N

HO2C HO2C

CO 2 H N H

HO2C

CO2H N

Photochemical degradation of enrofloxin.

the primary mode of degradation in the natural aquatic environment involves photolysis of the Fe complex (Lockhart and Blakeley 1975; Kari and Giger 1995). Its persistence is critically determined not only by the degree of insolation but also on the concentration of Fe in the environment, since complexes with other metals including Ca and Zn are relatively resistant to photolysis (Kari et al. 1995). The available evidence suggests that in contrast to nitrotriacetic acid (NTA) that is more readily biodegradable, EDTA is likely to be persistent except in environments in which concentrations of Fe greatly exceed those of other cations. The photolytic degradation of the fluoroquinolone enrofloxacin involves a number of reactions that produce 6-fluoro-7-amino-1-cyclopropylquinolone 2-carboxylic acid that is then degraded to CO2 via reactions involving fission of the benzenoid ring with loss of fluoride, dealkylation, and decarboxylation (Burhenne et al. 1997a,b) (Figure 1.9). Photolysis of the oxime group in the pyrazole miticide fenpyroximate resulted in the formation of two principal transformation products: the nitrile via an elimination reaction and the aldehyde by hydrolysis (Swanson et al. 1995). Photochemical transformation of pyrene in aqueous media produced the 1,6- and 1,8quinones as stable end products after initial formation of 1-hydroxypyrene (Sigman et al. 1998). The transformation of isoquinoline has been studied both under photochemical conditions with hydrogen peroxide, and in the dark with hydroxyl radicals (Beitz et al. 1998). The former resulted in fission of the pyridine ring with the formation of phthalic dialdehyde and phthalimide, whereas the major product from the latter reaction involved oxidation of the benzene ring with formation of the isoquinoline-5,8-quinone and a hydroxylated quinone. In the presence of both light and hydrogen peroxide, 2,4-dinitrotoluene is oxidized to the corresponding carboxylic acid; this is then decarboxylated to 1,3-dinitrobenzene, which is degraded further by hydroxylation and ring fission (Figure 1.10) (Ho 1986). Analogous reaction products were formed from 2,4,6-trinitrotoluene and hydroxylated to various nitrophenols and nitrocatechols before fission of the aromatic rings, and included the dimeric 2,2′-carboxy-3,3′,5,5′-tetranitroazoxybenzene (Godejohann et al. 1998). Nitrobenzene, 1-chloro-2,4-dinitrobenzene, 2,4-dinitrophenol, and 4-nitrophenol were degraded with the formation of formate, oxalate, and nitrate (Einschlag et al. 2002). It has been suggested that the photochemical reaction of pentachlorophenol in aqueous solution to produce octachlorodibenzo[1,4] dioxin and some of the heptachloro congener could account for the discrepancy between values for the emission of chlorinated dioxins and their deposition, which is significant for the octachloro congener (Baker and Hites 2000).

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8

Environmental Degradation and Transformation of Organic Chemicals CH3

CO2 H NO2

NO2

NO2

NO2

OH

OH NO2

NO2

NO2

O2 N

OH

CO2H CO2H

NO2

FIGURE 1.10

Me

N

FIGURE 1.11

NO 2

Photochemical transformation of 2,4-dinitrotoluene.

N Me

Me

N

N

Me

CO 2

MeNH 2 + CO 2

Photochemical transformation of paraquat.

CO 2 H Cl

Cl

FIGURE 1.12

N

CH 2 CO 2 H

CH 2

N H

N H

Photochemical transformation of diclofenac.

17. The psychopharmaceutical drug fluoxetine (Prozac) is degraded both directly and by the faster reaction with OH radicals (Lam et al. 2005). In both reactions, the ring bearing the CF 3 group was degraded in high yield to 4-(difluoromethylene)-cyclohexa-2,5-diene-1-one. 18. Whereas photolysis of 2- and 4-chlorophenols in aqueous solution produced catechol and hydroquinone, in ice the more toxic dimeric chlorinated dihydroxybiphenyls were formed (Bláha et al. 2004). 19. The photodegradation of the contact herbicide paraquat yielded many degradation products, but the major pathway produced 1,2,3,4-tetrahydro-1-ketopyrido[1,2-a]-5-pyrazinium that was further degraded to pyridine-2-carboxamide and pyridine-2-carboxylate (Figure 1.11) (Smith and Grove 1969). 20. Photodegradation of the nonsteroidal anti-inflammatory drug diclofenac produced carbazole-1-acetate as the major product (Figure 1.12) (Moore et al. 1990). In a lake under natural conditions, it was rapidly decomposed photochemically though none of the products produced in laboratory experiments could be detected (Buser et al. 1998). 21. Phototransformation of DDT in the presence of surfactants produced DDE by elimination and DDD by reductive dechlorination (Chu 1999).

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Abiotic Reactions

9

HYDROXYL RADICALS IN THE DESTRUCTION OF CONTAMINANTS The destruction of contaminants has directed attention to the use of hydroxyl radical–mediated reactions. These reactions should be viewed against those with hydroxyl radicals that occur in the atmosphere.

FENTON’S REAGENT Hydrogen peroxide in the presence of Fe2+ or Fe3+ (Fenton’s reagent) has been used in a range of configurations including irradiation, electrochemical, and both cathodic and anodic conditions (references in Wang et al. 2004). In all of these, the reaction involves hydroxyl radicals and has been studied particularly intensively for the destruction of agrochemicals including chlorinated phenoxyacetic acid (Sun and Pignatello 1993) and chloroacetanilide herbicides (Friedman et al. 2006). Systematic investigations have been carried out on the effect of pH, the molar ratio of H2O2/ substrate, and the possible complications resulting from the formation of iron complexes. Although this reaction may have limited environmental relevance except under rather special circumstances, it has been applied in combination with biological treatment of polycyclic aromatic hydrocarbons (PAHs) (Pradhan et al. 1997). Attention is drawn to it here since, under conditions where the concentration of oxidant is limiting, intermediates may be formed that are stable and that may possibly exert adverse environmental effects. Some examples that illustrate the formation of intermediates are given, although it should be emphasized that total destruction of the relevant xenobiotics under optimal conditions can be successfully accomplished. The structures of the products that are produced by the action of Fenton’s reagent on chlorobenzene are shown in Figure 1.13a (Sedlak and Andren 1991), on 2,4-dichlorophenoxyacetate in Figure 1.13b (Sun and Pignatello 1993), and on pentachlorophenol in Figure 1.13c (Fukushima and Tatsumi 2001). The UV-enhanced Fenton transformation of atrazine produced 2,4-diamino-6-hydroxy-1,3,5-triazine by a series of interacting reactions (Chan and Chu 2006). Whereas the degradation of azo dyes by Fenton’s reagent produced water- and CH2Cl2-soluble transformation products including nitrobenzene from Disperse Orange 3 O

Cl

(a)

Cl

Cl

OH ;

Cl

; O

O

CH 2 Cl

Cl O

Cl ;

(b)

Cl

O

O

Cl Cl

Cl

Cl

Cl

O

Cl

Cl

HO

Cl

(c)

Cl O

Cl

; Cl

FIGURE 1.13

Cl

Cl

Cl

O Cl

Cl Cl

Transformation products from (a) chlorobenzene, (b) 2,4-dichlorophenoxyacetate, (c) penta-

chlorophenol.

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that contains a nitro group, benzene was tentatively identified among volatile products from Solvent Yellow 14 (Spadaro et al. 1994).

POLYOXOMETALATES (HETEROPOLYACIDS) These are complexes formed between tungstates and molybdates, and silicate or phosphate, and have been used to generate hydroxyl radicals photochemically. The tungstates PW12O403− and SiW12O404− have been used most frequently. a. Degradation of 2,4,6-trichlorophenol by PW12O403− formed a number of products as intermediates, including 2,6-dichlorohydroquinone followed by fission of the ring to maleate, oxalate, acetate, and formate (Androulaki et al. 2000). b. Nonafluoropentanoic acid was decomposed in aqueous solution to fluoride and CO2 catalyzed by H3PW12O40 under UV-visible light radiation. The reaction was initiated by decarboxylation followed by a series of reactions involving oxidations (Hori et al. 2004b). c. The polyoxometalate PW12O403− was immobilized on an anion-exchange resin, and used to demonstrate the degradation in the presence of H2O2 of the phthalein dye rhodamine B to phthalate and a number of short-chain aliphatic mono- and dicarboxylates (Lei et al. 2005). They have also been used to bring about photochemical reduction of Hg2+ via Hg22+ to Hg0 (Troupis et al. 2005).

PHOTOLYTIC DEGRADATION ON TiO2 The mechanism involves photochemical production of a free electron in the conduction band (ecb−) and a corresponding hole (hvb+) in the valence band. Both of these produce H2O2 and thence hydroxyl radicals. a. In the presence of slurries of TiO2 that served as a photochemical sensitizer, methyl tert-butyl ether was photochemically decomposed at wavelengths <290 nm. The products were essentially the same as those produced by hydroxyl radicals under atmospheric conditions (Baretto et al. 1995): tert-butyl formate and tert-butanol were rapidly formed and further degraded to formate, acetone, acetate, and but-2-ene. b. The photocatalytic oxidation of various EDTA complexes has been examined (Madden et al. 1997). The rates and efficiencies were strongly dependent on the metal and the reactions are generally similar to those involved in electrochemical oxidation (Pakalapati et al. 1996). c. A number of products are formed from trichloroethene including tetrachloromethane, hexachloroethane, pentachloroethane, and tetrachlororethene, although the last two were shown to be degradable in separate experiments (Hung and Marinas 1997). In TiO2 slurries, the photochemical degradation of chloroform, bromoform, and tetrachloromethane involves initial formation of the trihalomethyl radicals. In the absence of oxygen, these are further decomposed via dihalocarbenes to CO. Dichlorocarbene was found as an intermediate in the degradation of trichloroacetate (Choi and Hoffmann 1997). d. The degradation of the herbicide 3-amino-1H-1,2,4-triazole (amitrole) produced a number of transformation products after fission of the triazole ring. These reacted together to form 2,4,6-trihydroxy-1,3,5-triazine that was a stable end product (Watanabe et al. 2005). e. The oxidation of tetrachlorobiphenyl congeners (23-34, 25-34, 345-4) with only one orthochlorine substituent was examined using nanostructured TiO2 immobilized on quartz beads. Initial reactions were hydroxylations by OH radicals that were followed by fission of a single ring to chlorobenzoates and low yields of succinate and glycolate (Nomiyama et al. 2005).

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OTHER PHOTOCHEMICALLY INDUCED REACTIONS 1. Two types of reactions are important in the photochemical transformation of PAHs, those with molecular oxygen and those involving cyclization. Illustrative examples are provided by the photooxidation of 7,12-dimethylbenz[a]anthracene (Lee and Harvey 1986) (Figure 1.14a) and benzo[a]pyrene (Lee-Ruff et al. 1986) (Figure 1.14b), and the cyclization of cis-stilbene (Figure 1.14c). 2. In nonaqueous solutions, two other types of reactions have been observed with polycyclic arenes: condensation via free-radical reactions and oxidative ring fission. a. Irradiation of benz[a]anthracene in benzene solutions in the presence of xanth-9-one or vanillin produced a number of transformation products tentatively identified as the result of oxidation and cleavage of ring A, ring C, ring D, and rings C and D, and rings B, C, and D, respectively (Jang and McDow 1997). b. 1-Nitropyrene is a widely distributed contaminant produced in the troposphere by reaction of nitrate radicals with pyrene, which is discussed later. A solution in benzene was photochemically transformed into 9-hydroxy-1-nitropyrene that is less mutagenic than its precursor (Koizumi et al. 1994). 3. The photochemical transformation of phenanthrene sorbed on silica gel (Barbas et al. 1996) resulted in a variety of products including cis-9,10-dihydrodihydroxyphenanthrene, phenanthrene-9,10-quinone, and a number of ring fission products including biphenyl2,2′-dicarboxaldehyde, naphthalene-1,2-dicarboxylic acid, and benzo[c]coumarin. 4. The products of the photooxidation of naphthylamines adsorbed on particles of silica and alumina were putatively less toxic than their precursors (Hasegawa et al. 1993). 5. Plant cuticles can play diverse roles. a. It has been suggested that photochemically induced reactions may take place between biocides and the biomolecules of plant cuticles. Laboratory experiments have examined addition reactions between DDT and methyl oleate, and were used to illustrate reactions, which resulted in the production of “bound” DDT residues (Figure 1.15) (Schwack 1988). O O

CH3

(a) CH3

CHO

(b) O O

O

(c)

FIGURE 1.14 Photooxygenation of (a) 7,12-dimethylbenz[a]anthracene, (b) benzo[a]pyrene, and (c) photocylization of cis-stilbene.

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Environmental Degradation and Transformation of Organic Chemicals Cl

Cl

CCl2 CH3(CH2)7 −CHCl−CH −(CH2)7 −CO2CH3

FIGURE 1.15 O

O

O

FIGURE 1.16

Product of reaction between DDT and methyl oleate. Cl

O

SO 2 Me

O

Cl

OH Cl

O

SO 2 Me

O

O

SO 2 Me

Photochemical transformation of sulcotrione.

b. The herbicide sulcotrione that is an alternative to atrazine is transformed photochemically on maize cuticular wax to a tetrahydroxanthone (ter Halle et al. 2006) (Figure 1.16). 6. There has been considerable interest in the photochemical decomposition of perfluorinated compounds that have become ubiquitous. a. Perfluorooctanoate can be successively degraded to C6, C5, C4, and C3 perfluoro acids, and to fluoride and CO2 (Hori et al. 2004a). The reaction is initiated by loss of CO2 with formation of the C7F15 radical that reacts with water to form C7F15OH. The following sequence of reactions is based on results from reactions in H218O: C7F15OH → C6F13COF + H+ + F− C6F13COF + H2O → C6F13CO2H + H+ + F− b. The decomposition of perfluorocarboxylic acids (C4 –C8) in the presence of the sulfate radical anion produced photochemically from persulfate (S2O82−) in the presence of oxygen resulted in the production of F−, CO2, and shorter-chain perfluoro fatty acids (Hori et al. 2005a). To facilitate mixing of the poorly water-soluble perfluorocarboxylic acids, the reaction has been carried out in a two-phase aqueous/liquid CO2 system (Hori et al. 2005b). High rates of degradation were observed for C9–C11 substrates, with the formation of fluoride, and only minor production of shorter-chain perfluorocarboxylic acids. 7. Photolysis of the potential fuel additive methylcyclopentadienyl manganese tricarbonyl resulted in the loss of CO and formation of a compound tentatively identified as the dicarbonyl (Wallington et al. 1999). In addition, there is concern over the particulates that are emitted during normal vehicle running. Speciation of the inorganic products has been carried out using manganese K-edge x-ray absorption fine structure (XANES) and showed that a number of products were formed including Mn3O4, MnSO4, and MnPO4 (Ressler et al. 2000).

THE ROLE OF HUMIC MATTER: SINGLET DIOXYGEN It is well established that important photochemical reactions are mediated by humic material in the aquatic environment (Zepp et al. 1981a,b), and that these are particularly significant for hydrophobic contaminants. Partial reductive dechlorination of the persistent insecticide mirex associated with

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humic matter has been reported (Burns et al. 1996, 1997; Lambrych and Hassett 2006), and it was shown that a kinetic model required differentiation of the bound and dissolved phases (Burns et al. 1996). The fundamental mechanism for the production of singlet dioxygen in aqueous solution was elucidated by Zepp et al. (1985). In the presence of the excited triplet states of chromophores such as humic acids, reaction with the triplet ground state of dioxygen (3O2) can produce singlet dioxygen (1O2) whose reaction with dienes has been traditionally used as a specific marker. Singlet 1O2 is highly reactive with a half-life of only 4 µs, but is significant in the aquatic environment in the neighborhood of the exciting chromophore. Using a trap-and-trigger chemiluminescent probe, it has been shown that within a volume containing the chromophore, high densities of 1O2 may be produced, and that this decreases during the transition to the true aquatic phase (Latch and McNeill 2006). Intrahumic reactions are therefore significant primarily for contaminants that are hydrophobic and that associate with the chromophoric groups of humic acids.

INTERACTIONS BETWEEN PHOTOCHEMICAL AND OTHER REACTIONS It has been shown that a combination of photolytic and biotic reactions can result in enhanced degradation of xenobiotics in municipal treatment systems, for example, of chlorophenols (Miller et al. 1988a) and benzo[a]pyrene (Miller et al. 1988b). Two examples illustrate the success of a combination of microbial and photochemical reactions in accomplishing the degradation of widely different xenobiotics in natural ecosystems. Both of them involved marine bacteria, and it therefore seems plausible to assume that such processes might be especially important in warm-water marine environments. a. The coupled degradation of pyridine dicarboxylates (Amador and Taylor 1990) b. The degradation of 3- and 4-trifluoromethylbenzoate: microbial transformation resulted in the formation of catechol intermediates that were converted into 7,7,7-trifluoro-hepta2,4-diene-6-one carboxylate, but this was subsequently degraded photochemically with the loss of fluoride (Taylor et al. 1993) (Figure 1.17). This degradation may be compared with the purely photochemical degradation of TFM that has already been noted, and contrasted with the resistance of trifluoromethylbenzoates to microbial degradation (Chapter 9, Part 3). Collectively, these examples illustrate the diversity of transformations of xenobiotics that are photochemically induced in aquatic and terrestrial systems. Photochemical reactions in the troposphere are extremely important in determining the fate and persistence of not only xenobiotics but also of naturally occurring compounds. A few illustrations are given as introduction: 1. The occurrence of C8 and C9 dicarboxylic acids in samples of atmospheric particles and in recent sediments (Stephanou 1992; Stephanou and Stratigakis 1993) has been attributed to photochemical degradation of unsaturated carboxylic acids that are widespread in almost all biota. 2. The formation of peroxyacetyl nitrate from isoprene (Grosjean et al. 1993a) and of peroxypropionyl nitrate (Grosjean et al. 1993b) from cis-3-hexen-1-ol that is derived from higher plants, illustrate important contributions to atmospheric degradation (Seefeld and Kerr 1997). CO 2 H

OH OH

CF 3

FIGURE 1.17

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CF 3

CO 2 H

O

CF 3

Microbial followed by photochemical degradation of 3-trifluorobenzoate.

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Environmental Degradation and Transformation of Organic Chemicals

3. Attention has been given to possible adverse effects of incorporating tert-butyl methyl ether into automobile fuels. tert-Butyl formate is an established product of photolysis, and it has been shown that photolysis in the presence of NO can produce the relatively stable tert-butoxyformyl peroxynitrate. This has a stability comparable to that of peroxyacetyl nitrate, and may therefore increase the potential for disseminating NOx (Kirchner et al. 1997).

REACTIONS IN THE TROPOSPHERE Although chemical transformations in the atmosphere may seem peripheral to this discussion, these reactions should be considered since their products may subsequently enter the aquatic and terrestrial environments: the persistence and the toxicity of these secondary products are therefore relevant to this discussion. Details of the relevant principles and details of the methodology are covered in a comprehensive treatise (Finlayson-Pitts and Pitts 1986), and reference should be made to a review on tropospheric air pollution (Finlayson-Pitts and Pitts 1997), and atmospheric aerosols (Andreae and Crutzen 1997). Attention has been directed to the potentially adverse reactions including ozone depletion induced by a range of anthropogenic compounds. The atmospheric chemistry of hydrofluorocarbons and hydrofluoroethenes is discussed by Wallington and Nielsen (2002), and organic bromine and iodine compounds by Orlando (2003). Although reactions with free radicals generally dominate, direct photolysis is important for alkyl iodides (Orlando 2003). There are a number of important reasons for discussing the reactions of organic compounds in the troposphere: 1. After emission, contaminants may be partitioned among the terrestrial, aquatic, and various atmospheric phases, and those of sufficient volatility or associated with particles may be transported over long distances. This is not a passive process, however, since important transformations may take place in the troposphere during transit so that attention should also be directed to their transformation products. 2. Considerable attention has been given to the persistence and fate of organic compounds in the troposphere, and this has been increasingly motivated by their possible role in the production of ozone by reactions involving NOx. 3. Concern has been expressed over the destruction of ozone in the stratosphere brought about by its reactions with chlorine atoms produced from chlorofluoroalkanes that are persistent in the troposphere, and that may contribute to radiatively active gases other than CO2. Reactions in the troposphere are mediated by reactions involving hydroxyl radicals produced photochemically during daylight, by nitrate radicals that are significant during the night (Platt et al. 1984), by ozone and, in some circumstances by O(3P). The overall reactions involved in the production of hydroxyl radicals are O3 + hν → O2 + O (1D); O(1D) + H2O → 2 OH O(1D) → O(3P); O(3P) + O2 → O3 Note that Roman capitals (S, P, D, etc.) are used for the states of atoms, and Greek capitals (∑, Π, ∆, etc.) for those of molecules, and that the ground state of O2 is a triplet O2(3∑). The reaction O3 + hν → O(1D) + O2(1∆) has an energy threshold at 310 nm, and the other possible reaction O3 + hν → O(1D) + O2(3∑) is formally forbidden by conservation of spin. Increasing evidence has, however, accumulated to show that the rate of production of O(1D), and therefore of hydroxyl

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radicals at wavelengths >310 nm is significant and that therefore, in contrast to previous assumptions, the latter reaction makes an important contribution (Ravishankara et al. 1998). Nitrate radicals are formed from NO that is produced during combustion processes and these are significant only during the night in the absence of photochemically produced OH radicals. They are formed by the reactions NO + O3 → NO2 + O2; NO2 + O3 → NO3 + O2 The concentrations of all these depend on local conditions, the time of day, and both altitude and latitude. Values of ca. 106 molecules/cm3 for OH, 108 –1010 molecules/cm3 for NO3, and ca. 1011 molecules/cm3 for ozone have been reported. Not all of these reactants are equally important, and the rates of reaction with a substrate vary considerably. Reactions with hydroxyl radicals are generally the most important, and some illustrative values are given for the rates of reaction (cm3/s/molecule) with hydroxyl radicals, nitrate radicals, and ozone (Atkinson 1990; summary of PAHs by Arey 1998) Hydroxyl Radicals 2.54 × 10–12 15.8 × 10–12 23.16 × 10–11

n-Butane Acetaldehyde Naphthalene

Nitrate Radicals 6.5 × 10–17 2.7 × 10–15 3.6 × 10–28 [NO2]

Ozone 9.8 × 10–24 < 10–20 < 2 × 10–19

SURVEY OF REACTIONS The major reactions carried out by hydroxyl and nitrate radicals may conveniently be represented for a primary alkane RH or a secondary alkane R2CH. In both, hydrogen abstraction is the initiating reaction. 1. Hydrogen abstraction RH + HO → R + H2O RH + NO3 → R + HNO3 2. Formation of alkylperoxy radicals R + O2 → R–O–O 3. Reactions of alkylperoxy radicals with NOx R–O–O + NO → R–O + NO2 R–O–O + NO → R–O–NO2 4. Reactions of alkyloxy radicals R2CH–O + O2 → R2CO + HO2 R2CH–O + NO → R2CH–O–NO2 → R2CO + HNO

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The concentration of NO determines the relative importance of reaction 3, and the formation of NO2 is particularly significant since this is readily photolyzed to produce O(3P) that reacts with oxygen to produce ozone. This alkane–NOx reaction may produce O3 at the troposphere–stratosphere interface: NO2 → NO + O(3P); O(3P) + O2 → O3 This is the main reaction for the formation of ozone although, under equilibrium conditions, the concentrations of NO2, NO, and O3 are interdependent and no net synthesis of O3 occurs. When, however, the equilibrium is disturbed and NO is removed by reactions with alkylperoxy radicals (reactions 1+2+3), synthesis of O3 may take place. RH + OH → R + H2O; R + O2 → RO2; RO2 + NO → RO + NO2 The extent to which this occurs depends on a number of issues (Finlayson-Pitts and Pitts 1997), including the reactivity of the hydrocarbon that is itself a function of many factors. It has been proposed that the possibility of ozone formation is best described by a reactivity index of incremental hydrocarbon reactivity (Carter and Atkinson 1987, 1989) that combines the rate of formation of O3 with that of the reduction in the concentration of NO. The method has been applied, for example, to oxygenate additives to automobile fuel (Japar et al. 1991), while both anthropogenic compounds and naturally occurring hydrocarbons may be reactive. Clearly, whether or not ozone is formed depends also on the rate at which, for example, unsaturated hydrocarbons react with it. Rates of reactions of ozone with alkanes are, as noted above, much slower than for reaction with OH radicals, and reactions with ozone are of the greatest significance with unsaturated aliphatic compounds. The pathways plausibly follow those involved in chemical ozonization (Hudlický 1990). The kinetics of the various reactions have been explored in detail using large-volume chambers that can be used to simulate reactions in the troposphere. They have frequently used hydroxyl radicals formed by photolysis of methyl (or ethyl) nitrite, with the addition of NO to inhibit photolysis of NO2. This would result in the formation of O(3P) atoms, and subsequent reaction with O2 would produce ozone, and hence NO3 radicals from NO2. Nitrate radicals are produced by the thermal decomposition of N2O5, and in experiments with O3, a scavenger for hydroxyl radicals is added. Details of the different experimental procedures for the measurement of absolute and relative rates have been summarized, and attention drawn to the often considerable spread of values for experiments carried out at room temperature (~298 K) (Atkinson 1986). It should be emphasized that in the real troposphere, both the rates—and possibly the products—of transformation will be determined by seasonal differences both in temperature and the intensity of solar radiation. These are determined both by latitude and altitude. The kinetics of the reactions of several xenobiotics with hydroxyl or nitrate radicals have been examined under simulated atmospheric conditions and include those given in Table 1.1. For polychlorinated biphenyls (PCBs), rate constants were highly dependent on the number of chlorine atoms, and calculated atmospheric lifetimes varied from 2 d for 3-chlorobiphenyl to 34 d for 236-25 pentachlorobiphenyl (Anderson and Hites 1996). It was estimated that loss by hydroxylation in the atmosphere was a primary process for the removal of PCBs from the environment. It was later shown that the products were chlorinated benzoic acids produced by initial reaction with a hydroxyl radical at the 1-position followed by transannular dioxygenation at the 2- and 5-positions followed by ring fission (Brubaker and Hites 1998). Reactions of hydroxyl radicals with polychlorinated dibenzo[1,4]dioxins and dibenzofurans also play an important role for their removal from the atmosphere (Brubaker and Hites 1997). The gas phase and the particulate phase are in equilibrium, and the results show that gas-phase reactions with hydroxyl radicals are important for the

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TABLE 1.1 Degradation of Contaminants under Simulated Atmospheric Conditions Xenobiotic Aliphatic and aromatic hydrocarbons Substituted monocyclic aromatic compounds Terpenes Amines Heteroarenes Chlorinated aromatic hydrocarbons Volatile methyl–silicon compounds Methylcyclopentadienyl manganese (CO)3

Reference Tuazon et al. (1986) Tuazon et al. (1986) Atkinson et al. (1985a) Atkinson et al. (1987) Atkinson et al. (1985b) Kwok et al. (1995) Tuazon et al. (2000) Wallington et al. (1999)

compounds with fewer numbers of chlorine substituents, whereas for those with higher numbers of substituents particle-phase removal is significant. Considerable attention has been directed in determining the products from reactions of aliphatic hydrocarbons, aromatic compounds, and unsaturated compounds including biogenic terpenes that exhibit appreciable volatility. These studies have been conducted both in simulation chambers and using natural sunlight in the presence of NO.

SURVEY OF REACTANTS Aliphatic Hydrocarbons A range of transformation products has been identified from simulated reactions of alkanes. These include alkyl nitrates by reactions that have already been given, but also include a range of hydroxycarbonyls that are summarized in Table 1.2 (Reisen et al. 2005). Aromatic Hydrocarbons Fission of the rings of aromatic hydrocarbons may take place. For example, o-xylene forms diacetyl, methylglyoxal, and glyoxal (Tuazon et al. 1986), which are also produced by ozonolysis (Levine and Cole 1932), while naphthalene forms 2-formylcinnamaldehyde (Arey 1998). The photooxidation of alkyl benzenes that are atmospheric contaminants with high volatility has been studied in detail and the reaction pathways have been delineated (Yu et al. 1997). Products from alkyl benzenes included both those with the ring intact, such as aromatic aldehydes and quinones, together with a wide range of aliphatic compounds containing alcohol, ketone, and epoxy functional groups resulting from ring fission. The significance of epoxide intermediates (Yu and Jeffries 1997) is noted in the next section, and attention is drawn later to the important reactions of arenes that result in the production of nitroarenes. Biogenic Terpenes Monoterpenes are appreciably volatile and are produced in substantial quantities by a range of higher plants and trees. Only summary of examples are given here. 1. Photochemical reactions of isoprene (references in Grosjean et al. 1993a). 2. Products from reaction of α-pinene with ozone that produced a range of cyclobutane carboxylic acids (Kamens et al. 1999). 3. Rapid reactions of linalool with OH radicals, NO3 radicals, and ozone in which the major products were acetone and 5-ethenyldihydro-5-methyl-2(3H)-furanone (Shu et al. 1997).

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TABLE 1.2 Products from Simulated Reactions of Alkanes Alkane n-Pentane n-Hexane

n-Heptane

n-Octane

Product

Alkane

Product

5-Hydroxypentan-2-one 5-Hydroxypentanal 5-Hydroxyhexan-2-one 6-Hydroxyhexan-3-one 4-Hydroxyhexanal 5-Hydroxyheptan-2-one 6-Hydroxyheptan-3-one 1-Hydroxyheptan-4-one 4-Hydroxyheptanal 5-Hydroxyoctan-2-one 6-Hydroxyoctan-3-one 7-Hydroxyoctan-4-one 4-Hydroxyoctanal

n-Heptane

5-Hydroxyheptan-2-one 6-Hydroxyheptan-3-one 1-Hydroxyheptan-4-one 4-Hydroxyheptanal 5-Hydroxyoctan-2-one 6-Hydroxyoctan-3-one 7-Hydroxyoctan-4-one 4-Hydroxyoctanal

n-Octane

4. Plant metabolite cis-hex-3-ene-1-ol that is the precursor of peroxypropionyl nitrate (Grosjean et al. 1993b) analogous to peroxyacetyl nitrate. 5. Degradation of many terpenes has been examined including β-pinene, d-limonene, and trans-caryophyllene (Grosjean et al. 1993c). 6. Products formed by reaction of NO3 radicals with α-pinene have been identified and include pinane epoxide, 2-hydroxypinane-3-nitrate, and 3-ketopinan-2-nitrate formed by reactions at the double bond, and pinonaldehyde that is produced by ring fission between C-2 and C-3 (Wängberg et al. 1997). These reactions should be viewed in the general context of “odd nitrogen” to which alkyl nitrates belong (Schneider et al. 1998). 7. Gas-phase products from the reactions of ozone with the monoterpenes (–)-β-pinene and (+)-sabinene included the ketones formed by oxidative fission of the exocyclic C=C bonds as well as ozonides from the addition of ozone to this bond (Griesbaum et al. 1998).

REENTRY OF TROPOSPHERIC TRANSFORMATION PRODUCTS An important aspect of atmospheric reactions is the possibility that tropospheric transformation products subsequently enter aquatic and terrestrial ecosystems through precipitation or by particle deposition.

HALOGENATED ALKANES AND ALKENES The stability of perchlorofluoroalkanes is due to the absence of hydrogen atoms that may be abstracted by reaction with hydroxyl radicals. Attention has therefore been directed to chlorofluoroalkanes containing at least one hydrogen atom (Hayman and Derwent 1997). Considerable effort has also been directed to the reactions of chloroalkanes and chloroalkenes, and this deserves a rather more detailed examination in the light of interest in the products that are formed. a. There has been concern over the fate of halogenated aliphatic compounds in the atmosphere, and a single illustration of the diverse consequences is noted here. The initial

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reaction of 1,1,1-trichloroethane with hydroxyl radicals produces the Cl3C–CH2 radical by abstraction of H and then undergoes a complex series of reactions including the following: Cl3C–CH2 + O2 → Cl3C–CH2O2 2Cl3C–CH2O2 → 2Cl3C–CH2O Cl3C–CH2O + O2 → Cl3C–CHO In addition, the alkoxy radical Cl3C–CH2O produces highly reactive COCl2 (phosgene) (Platz et al. 1995; Nelson et al. 1990) that has been identified in atmospheric samples and was attributed to the transformation of gem-dichloro aliphatic compounds (Grosjean 1991). b. The atmospheric degradation of tetrachloroethene is initiated by reaction with Cl radicals and produces trichloroacetyl chloride as the primary intermediate which is formed followed by the following reactions (Franklin 1994): Cl2C=CCl2 + Cl → Cl3C–CCl2 Cl3C–CCl2 + O2 → Cl3C–CCl2O2 Cl3C–CCl2O2 + NO → Cl3C–CCl2O + NO2 Cl3C–CCl2O → Cl3C–COCl + Cl → COCl2 + CCl3 CCl3 + O2 + NO → COCl2 + NO2 + Cl An overview of the reactions involving trihalomethanes (haloforms) CHXYZ, where X, Y, and Z are halogen atoms, has been given in the context of ozone depletion (Hayman and Derwent 1997). Interest in the formation of trichloroacetaldehyde formed from trichloroethane and tetrachloroethene is heightened by the phytotoxicity of trichloroacetic acid (Frank et al. 1994), and by its occurrence in rainwater that seems to be a major source of this contaminant (Müller et al. 1996). The situation in Japan seems, however, to underscore the possible significance of other sources including chlorinated wastewater (Hashimoto et al. 1998). Whereas there is no doubt about the occurrence of trichloroacetic acid in rainwater (Stidson et al. 2004), its major source is unresolved since questions remain on the rate of hydrolysis of trichloroacetaldehyde (Jordan et al. 1999). Low concentrations of trifluoroacetate have been found in lakes in California and Nevada (Wujcik et al. 1998). It is formed by atmospheric reactions from 1,1,1,2-tetrafluoroethane and from the chlorofluorocarbon replacement compound CF3–CH2F (HFC-134a) in an estimated yield of 7–20% (Wallington et al. 1996). CF3OH that is formed from CF3 in the stratosphere is apparently a sink for its oxidation products (Wallington and Schneider 1994). CF3 –CH2F + OH → CF3–CHF CF3–CHF + O2 → CF3 –CHFO2 CF3–CHFO2 + NO → CF3–CHFO + NO2 CF3–CHFO + O2 → CF3 –COF CF3 –CHFO → CF3 + H–COF

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It has been suggested that an alternative source of trifluoroacetic acid in the environment may be provided by the aqueous photolysis of TFM and related trifluoromethylphenols (Ellis and Mabury 2000). Although trifluoroacetate is accumulated by a range of biota through incorporation into biomolecules (Standley and Bott 1998), unlike trichloroacetate, it is only weakly phytotoxic, and there is no evidence for its inhibitory effect on methanogenesis (Emptage et al. 1997). Attention has been directed to the transformation of perfluorotelomer alcohols CF3(CF2)7–CH2CH2OH that could be the source of the ubiquitous perfluorinated carboxylic acids (PFCAs), and that have been produced during simulated atmospheric reactions with chlorine radicals as surrogate for hydroxyl radicals (Ellis et al. 2004).

ARENES AND NITROARENES The transformation of arenes in the troposphere has been discussed in detail (Arey 1998). Their destruction can be mediated by reaction with hydroxyl radicals, and from naphthalene a wide range of compounds is produced, including 1- and 2-naphthols, 2-formylcinnamaldehyde, phthalic anhydride, and with less certainty 1,4-naphthoquinone and 2,3-epoxynaphthoquinone. Both 1- and 2-nitronaphthalene were formed through the intervention of NO2 (Bunce et al. 1997). Attention has also been directed to the composition of secondary organic aerosols from the photooxidation of monocyclic aromatic hydrocarbons in the presence of NOx (Forstner et al. 1997); the main products from a range of alkylated aromatics were 2,5-furandione and the 3-methyl and 3-ethyl congeners. Considerable attention has been directed to the formation of nitroarenes that may be formed by several mechanisms: (a) initial reaction with hydroxyl radicals followed by reactions with nitrate radicals or NO2 and (b) direct reaction with nitrate radicals. The first is important for arenes in the troposphere, whereas the second is a thermal reaction that occurs during combustion of arenes. The kinetics of formation of nitroarenes by gas-phase reaction with N2O5 has been examined for naphthalene (Pitts et al. 1985a) and methylnaphthalenes (Zielinska et al. 1989); biphenyl (Atkinson et al. 1987b,c); acephenanthrylene (Zielinska et al. 1988); and for adsorbed pyrene (Pitts et al. 1985b). Both 1- and 2-nitronaphthalene were formed through OH radical–initiated reactions with naphthalene by the intervention of NO2 (Bunce et al. 1997). The major product from the first group of reactions is 2-nitronaphthalene, and a number of other nitroarenes have been identified including nitropyrene and nitrofluoranthenes (Arey 1998). The tentative identification of hydroxylated nitroarenes in air particulate samples (Nishioka et al. 1988) is consistent with the operation of this dual mechanism. Reaction of methyl arenes with nitrate radicals in the gas phase gives rise to a number of products. From toluene, the major product was benzaldehyde with lesser amounts of 2-nitrotoluene > benzyl alcohol nitrate > 4-nitrotoluene > 3-nitrotoluene (Chlodini et al. 1993). An important example is the formation of the mutagenic 2-nitro- and 6-nitro-6H-dibenzo[b,d]pyran-6-ones (Figure 1.18) from the oxidation of phenanthrene in the presence of NOx and methyl nitrite as a source of hydroxyl radicals (Helmig et al. 1992a). These compounds have been identified in samples of ambient air (Helmig et al. 1992b), and analogous compounds from pyrene have been tentatively identified (Sasaki et al. 1995). These compounds add further examples to the list of mononitroarenes that already include 2-nitropyrene and 2-nitrofluoranthene, and it appears plausible to suggest that NO 2

O O

FIGURE 1.18

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Product from the photochemical reaction of phenanthrene and NOx.

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21

comparable reactions are involved in the formation of the 1,6- and 1,8-dinitroarenes that have been identified in diesel exhaust. 3-Nitrobenzanthrone that is formally analogous to the dibenzopyrones noted earlier has also been identified in diesel exhaust and is also highly mutagenic to Salmonella typhimurium strain TA 98 (Enya et al. 1997). Many nitroarenes including 2-nitronaphthalene are direct-acting, frameshift mutagens in the Ames test (Rosenkranz and Mermelstein 1983). Although the mechanism has not been finally resolved, it appears to involve metabolic participation of the test organisms, for example, by reduction. In addition, nitroarenes may be reduced microbiologically to the corresponding amino compounds, and in terrestrial and aquatic systems these may have undesirable properties: (a) some including 2-aminonaphthalene are carcinogenic to mammals and (b) they react with components of humic and fulvic acids, which makes them more recalcitrant to degradation and therefore more persistent in ecosystems. Although nitrofluoranthenes and nitropyrenes are established atmospheric products from the nitration of the respective PAHs, their direct mutagenic activity is unable to account for the total mutagenic activity of airborne samples. It has been shown that the concentrations of 1- and 2-nitrotriphenylenes were higher than those of 1-nitropyrene, and that 1-nitrotriphenylene was a highly active direct-acting mutagen in S. typhimurium strains TA98 and YG 1024 (Ishii et al. 2000). Nitrated phenols have been found in rainwater, and are secondary products from atmospheric oxidation of benzene and alkylated toluenes followed by nitration (Kohler and Heeb 2003). Plausible mechanisms for the formation of 2,4-dinitrophenol from 2- and 4-nitropohenol in the atmospheric liquid phase have been proposed (Vione et al. 2005), and include nitration by NO2 or NO3 radicals. A wide range of azaarenes is produced during combustion (Herod 1998), and these may enter the troposphere, with formation of the corresponding nitro derivatives.

ALKYLATED ARENES The products from the oxidation of alkylbenzenes under simulated atmospheric conditions have been examined. Both ring epoxides that were highly functionalized, and aliphatic epoxides from ring fission were tentatively identified (Yu and Jeffries 1997). Formation of the epoxides, many of which are mutagenic, may cause concern over the transformation products of monocyclic aromatic hydrocarbons in the atmosphere. Toluene is an important atmospheric contaminant and its photochemistry has been explored using EUROPHORE, which is described in Chapter 5 (Klotz et al. 1998). This used natural illumination in a large-volume chamber equipped with differential optical spectroscopy (path length 326.8 m) and Fourier transform spectroscopy (path length 128 m) to minimize artifacts from sampling or treatment for analysis. Concentrations of NOx resembled those found in the atmosphere, and concentrations of ozone were low at the beginning of the experiment. The major reaction involved OH radicals that initiated two reactions: (a) removal of a hydrogen atom from the methyl group to form benzaldehyde and (b) formation of a toluene-hydroxyl adduct (methyl-hydroxy-cyclohexadienyl radical) that reacted with oxygen to produce o-, m-, and p-cresol that reacted further with nitrate radicals.

SULFIDES AND DISULFIDES Sulfides and disulfides can be produced by bacterial reactions in the marine environment. 2-Dimethylthiopropionic acid is produced by algae and by the marsh grass Spartina alternifolia, and may then be metabolized in sediment slurries under anoxic conditions to dimethyl sulfide (Kiene and Taylor 1988), and by aerobic bacteria to methyl sulfide (Taylor and Gilchrist 1991). Further details are given in Chapter 11, Part 2. Methyl sulfide can also be produced by biological methylation of sulfide itself (HS−). Carbon radicals are not the initial atmospheric products from organic sulfides and disulfides, and the reactions also provide an example in which the rates of reaction with nitrate

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radicals exceed those with hydroxyl radicals. Dimethyl sulfides—and possibly methyl sulfide as well—are oxidized in the troposphere to sulfur dioxide and methanesulfonic acids: CH3 –SH → CH3–SO3H CH3–S–S–CH3 → CH3–SO3H + CH3 –SO CH3–SO → CH3 + SO2 It has been suggested that these compounds may play a critical role in promoting cloud formation (Charleson et al. 1987), so that the long-term effect of the biosynthesis of methyl sulfides on climate alteration may be important. And yet, at first glance this seems far removed from the production of an osmolyte by marine algae and its metabolism in aquatic systems. The occurrence of methyl sulfates in atmospheric samples (Eatough et al. 1986) should be noted, although the mechanism of its formation appears not to have been established. These reactions provide a good example of the long chain of events that may bring about environmental effects through the subtle interaction of biotic and abiotic reactions in both the aquatic and atmospheric environments. Appreciation of interactive processes that have been outlined has been able to illuminate discussion of mechanisms for reactions as diverse as the acidification of water masses, climate alteration, ozone formation and destruction, and the possible environmental roles of trichloroacetic acid and nitroarenes.

CHEMICALLY MEDIATED TRANSFORMATIONS Only a limited number of the plethora of known chemical reactions has been observed in transforming xenobiotics in the environment. An attempt is made merely to present a classification of the reactions that take place with illustrations of their occurrence.

HYDROLYSIS Organic compounds containing carbonyl groups flanked by alkoxy groups (esters) or by amino- or substituted amino groups (amides, carbamates, and ureas) may be hydrolyzed by purely abiotic reactions under appropriate conditions of pH. The generally high pH of seawater (ca. 8.2) suggests that chemical hydrolysis may be important in this environment. In contrast, although very few natural aquatic ecosystems have pH values sufficiently low for acidic hydrolysis to be of major importance, this could be important in terrestrial systems when the pH is lowered by fermentation of organic substrates. It is therefore important to distinguish between alkaline or neutral, and acidic hydrolytic mechanisms. It should also be appreciated that both hydrolytic and photolytic mechanisms may operate simultaneously and that the products of these reactions may not necessarily be identical. It has already been noted that substantial numbers of important agrochemicals contain carbonyl groups, so that abiotic hydrolysis may be the primary reaction in their transformation: the example of carbaryl has already been cited (Wolfe et al. 1978). The same general principles may be extended to phosphate and thiophosphate esters, although in these cases it is important to bear in mind the stability to hydrolysis of primary and secondary phosphate esters under the neutral or alkaline conditions that prevail in most natural ecosystems. In contrast, sulfate esters and sulfamides are generally quite resistant to chemical hydrolysis except under rather drastic conditions, and microbial sulfatases and sulfamidases generally mediate their hydrolysis. Some examples illustrate the range of hydrolytic reactions that involve diverse agrochemicals that may enter aquatic systems by leaching from the soil.

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23

1. The cyclic sulfite of α- and β-endosulfan (Singh et al. 1991). 2. The carbamate phenmedipham that results in the intermediate formation of m-tolyl isocyanate (Figure 1.19) (Bergon et al. 1985). 3. 2-(Thiocyanomethylthio)benzthiazole initially forms 2-thiobenzthiazole, which is transient and then rapidly degrades photochemically to benzothiazole and 2-hydroxybenzothiazole (Brownlee et al. 1992). 4. Aldicarb undergoes simple hydrolysis at pH values above 7, whereas at pH values <5 an elimination reaction intervenes (Figure 1.20) (Bank and Tyrrell 1984). 5. The sulfonyl urea sulfometuron methyl is stable at neutral or alkaline pH values, but is hydrolyzed at pH 5 to methyl 2-aminosulfonylbenzoate that is cyclized to saccharin (Figure 1.21) (Harvey et al. 1985). The original compound is completely degraded to CO2 by photolysis. 6. The pyrethroid insecticides fenvalerate and cypermethrin are hydrolyzed under alkaline conditions at low substrate concentrations, but at higher concentrations the initially formed 3-phenoxybenzaldehyde reacts further with the substrate to form dimeric compounds (Figure 1.22) (Camilleri 1984). 7. The sulfonyl urea herbicide rimsulfuron is degraded increasingly rapidly in the pH range 5–9. The main degradation pathway is by rearrangement of the sulfonyl urea group followed by hydrolysis (Schneiders et al. 1993) (Figure 1.23). OH O−CO−NH

CH 3

CH 3

N

C

O CH 3

NH 2

+ NH−CO2CH3 NH−CO2CH3

FIGURE 1.19

Hydrolysis of phenmedipham. CH 3

SCH 3 C

CH 3

CH 3

CH

N−O−CONHCH3

CH 3

SCH 3 C

CH 3

SCH 3 C

C

CH 3

N

CH

NOH

+ CH 3 N

FIGURE 1.20

C

O

CH 3NH 2 + CH3NH−CO−NHCH3

Hydrolysis of aldicarb.

CO2CH3

CH3

CO2CH3

O

SO2NH2

SO2

NH

N SO2NH−CO−NH N CH 3

FIGURE 1.21

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Hydrolysis of sulfometuron.

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Environmental Degradation and Transformation of Organic Chemicals CN O

CH−O−COR

O RCO 2H

CHO

+

O−COR O

FIGURE 1.22

CO−CH

O

Hydrolysis of the pyrethroid insecticides fenvalerate and cypermethrin.

SO2Et SO2−NH

N

SO 2 Et

SO 2 Et N

CO−NH2

N

CO N

HN N MeO

FIGURE 1.23

N

N

MeO

N

NH

N

OMe

MeO

N OMe

OMe

Hydrolysis of rimsulfuron.

SO2 CH 3 O

SO2 CH 3 O

SO2 CH 3

O

CO2 H N F3 C

FIGURE 1.24

O

F3 C

C N

F3 C

Degradation of isoxaflutole.

8. The thiophosphate phorate is degraded in aqueous solutions at pH 8.5 to yield diethyl sulfide and formaldehyde, which are formed by nucleophilic attack either at the P = S atom or the methylene dithioketal carbon atom (Hong and Pehkonen 1998). 9. Hydrolysis of the isoxazole herbicide isoxaflutole rapidly produces a biologically active diketonitrile that is further degraded to products, which are partly bound to soil components and partly degraded to CO2 (Beltran et al. 2000; Lin et al. 2002; Taylor-Lovell et al. 2002) (Figure 1.24). 10. In alkaline solution, 3-bromo-2,2-bis(bromomethyl)propanol undergoes successive loss of bromide to produce 2,6-dioxaspiro(3,3)-heptane (Ezra et al. 2005) (Figure 1.25). 11. The reaction of brominated diphenyl ethers with methoxide in dimethyl formamide has been examined, and suggested as a ranking of their susceptibility to hydrolytic reactions under natural conditions (Rahm et al. 2005). The nature of the products was not apparently systematically examined.

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Abiotic Reactions

25 Br Br

Br

Br

Br

OH O

FIGURE 1.25

O

O

Degradation of 3-bromo-2,2-bis(bromomethyl)propanol.

Three important comments should be added: 1. Abiotic hydrolysis generally accomplishes only a single step in the ultimate degradation of the compounds that have been used for illustration. The intervention of subsequent biotic reactions is therefore almost invariably necessary for their complete mineralization. 2. The operation of these hydrolytic reactions is independent of the oxygen concentration of the system so that—in contrast to biotic degradation and transformation—these reactions may occur effectively under both aerobic and anaerobic conditions. 3. Rates of hydrolysis may be influenced by the presence of dissolved organic carbon, or organic components of soil and sediment. The magnitude of the effect is determined by the structure of the compound and by the kinetics of its association with these components. For example, whereas the neutral hydrolysis of chlorpyrifos was unaffected by sorption to sediments, the rate of alkaline hydrolysis was considerably slower (Macalady and Wolf 1985); humic acid also reduced the rate of alkaline hydrolysis of 1-octyl 2,4-dichlorophenoxyacetate (Perdue and Wolfe 1982). Conversely, sediment sorption had no effect on the neutral hydrolysis of 4-chlorostilbene oxide, although the rate below pH 5 where acid hydrolysis dominates was reduced (Metwally and Wolfe 1990).

REDUCTIVE DISPLACEMENT: DEHALOGENATION AND DESULFURIZATION There has been considerable interest in the abiotic dechlorination of chlorinated ethenes at contaminated sites. Reductive dehalogenation has therefore been examined using a range of reductants, many of them involving reduced complexes of porphyrins or corrins.

REDUCTIVE DEHALOGENATION ZERO-VALENT METALS Zero-valent iron has been used for diverse dehalogenations: a. The reductive dechlorination of DDT with zero-valent iron in an anaerobic aqueous medium produced DDD by reductive dechlorination, and this was further transformed at higher concentrations of Fe (Sayles et al. 1997). b. Zero-valent iron was able to degrade octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine (HMX) to hydrazine and formaldehyde, putatively via the intermediate methylendinitramine (Monteil-Rivera et al. 2005). c. Zero-valent metals in subcritical water have been applied to the dechlorination of PCBs (Kubátova et al. 2003; Yak et al. 1999), and is noted again later. Attention has been particularly devoted to the application of nanoscale particles of zero-valent iron, and a range of halogenated aliphatic compounds has been examined.

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1. For trichloroethene (TCE), the stoichiometric amount of iron and the effect of different preparations determine the outcome of the several competing reactions. Coupling products such as butenes, acetylene and its reduction products ethene and ethane, and products with five or six carbon atoms were formed (Liu et al. 2005). Although a field-scale application successfully lowered the concentration of TCE, there was evidence for the formation of the undesirable cis-1,2-dichloroethene and 1-chloroethene (vinyl chloride) in the groundwater (Quinn et al. 2005). 2. A range of chlorinated ethanes has been examined and several transformations were observed (Song and Carraway 2005). a. Combined β-elimination and reductive dechlorination: hexachloroethane → tetrachloroethene pentachloroethane → tetrachloroethene and trichloroethene 1,1,2,2-tetrachloroethane → cis- and trans-dichloroethene b. Reductive dechlorination: 1,1,1-trichloroethane → 1,1-dichloroethane; c. α-Elimination: 1,1-dichloroethane → ethane. 3. Experiments have been carried out to compare nanoscale catalysts composed of Fe-, Ni-, and Co-complexes of several porphyrins or cyanocobalamin (Dror et al. 2005). A cobalt–porphyrin complex and cyanocobalamin in the presence of Ti(III)citrate reduced the initial concentrations of tetrachloromethane and tetrachloroethene by ≈99.5%, and the porphyrin was equally effective with trichloroethene. The advantage of using heterogeneous catalysts was shown by experiments in repetitive cycling of tetrachloromethane. Zero-valent metals degrade vicinal dichlorides such as tetrachloroethene by α-elimination to produce dichloroacetylene and finally acetylene (Roberts et al. 1996).

CHROMOUS CHLORIDE 1. Cr(II) has been used to bring about dehalogenation of alkyl halides involving the production of alkyl radicals, and details have been provided in a substantive review (Castro 1998). The ease of reduction is generally: iodides > bromides > chlorides, while tertiary halides are the most reactive and primary halides the least (Castro and Kray 1963, 1966). 2. The stereospecificity of dehalogenation of vicinal dibromides to olefins was examined for reducing agents including Cr(II), iodide, and Fe0 (Totten et al. 2001). For dibromostilbene, the (E)-stilbene represented >70% of the total olefin that was produced, and for threodibromopentane reduction by Cr(II) produced ca. 70% of (E)-pent-2-ene, whereas values for iodide and Fe0 were <5% of this.

PORPHYRINS AND CORRINS 1. Attention has been directed to the role of corrins and porphyrins in the absence of biological systems. A range of structurally diverse compounds can be dechlorinated: these include DDT (Zoro et al. 1974); lindane (Marks et al. 1989); mirex (Holmstead 1976); C1 chloroalkanes (Krone et al. 1989); C2 chloroalkenes (Gantzer and Wackett 1991), and C2 chloroalkanes (Schanke and Wackett 1992). Detailed mechanistic examination of the dehydrochlorination of pentachloroethane to tetrachloroethene reveals, however, the potential complexity of this reaction, and the possibly significant role of pentachloroethane in the abiotic transformation of hexachloroethane (Roberts and Gschwend 1991). 2. Considerable attention has been directed to dehalogenation mediated by corrinoids and porphyrins in the presence of a chemical reductant (references in Gantzer and Wackett 1991; Glod et al. 1997; Workman et al. 1997). Illustrations are provided by the dechlorination and elimination reactions carried out by titanium(III) citrate and hydroxocobalamin (Bosma et al. 1988; Glod et al. 1997). The involvement of corrinoids and porphyrins is consistent with the occurrence of analogous mechanisms for biological reactions that

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3.

4.

5.

6.

27

are generally transformations involving dehydrochlorination (elimination) and reductive dechlorination. a. Chlorofluoroalkanes—interest in their adverse environmental effects has stimulated interest in their anaerobic transformation. Although this is probably mediated by abiotic reactions involving porphyrins (Lovley and Woodward 1992; Lesage et al. 1992), it should be noted that the C–F bond was apparently retained in the products (Lesage et al. 1992). b. Dechlorination and elimination reactions of hexachlorobuta-1,3-diene took place in the presence of titanium(III) citrate and hydroxocobalamin (Bosma et al. 1994). Hexachlorobuta-1,3-diene was dechlorinated to the pentachloro compound and, by dechlorination and elimination, successively to trichlorobut-1-ene-3-yne (probably the 1,2,2-trichloro isomer), and but-1-ene-2-yne. The specificity of corrins and porphyrins is of particular interest since it seems to be significantly less than that of the enzymes generally implicated in microbial dechlorination. In aqueous solution containing cobalamin and titanium(III) as reductant, tetrachloroethene was dehalogenated to the trichloroethene radical, and thence to the dichlorovinyl radical, chloroacetylene, and acetylene (Glod et al. 1997). c. Reductive dechlorination of the α-, β-, γ-, and δ-isomers of hexachlorocyclohexane was examined under several conditions. The most effective was the combination of titanium(III) citrate and hydroxocobalamin, when all the isomers were degraded in the order of rates γ > α > δ > β. Tetrachlorocyclohexene was formed as a transient intermediate with chlorobenzene as the ultimate product (Rodríguez-Garrido et al. 2004). An additional aspect of these dehalogenations that elucidates the role of vitamin B12 is provided by experiments with Shewanella alga strain BrY (Workman et al. 1997). This organism carries out reduction of Fe(III) and Co(III) during growth with lactate and H2, and was used to reduce vitamin B12a anaerobically in the presence of an electron donor. The biologically reduced vitamin B12 was then able to transform tetrachloromethane to CO. The reactions of alkyl halides with Fe(II) deuteroporphyrin IX have been examined (Wade and Castro 1973). Three classes of reaction were observed (i) hydrogenolysis, (ii) elimination to alkenes, and (iii) coupling of alkyl free radicals. Further discussion has been given in Castro (1998). Experiments have been carried out to mimic the reactions of model systems for coenzyme F430 that is involved in the terminal step in the biosynthesis of methane, and that is able to dechlorinate CCl4 successively to CHCl3 and CH2Cl2 (Krone et al. 1989). Nickel(I) isobacteriochlorin anion was generated electrolytically and used to examine the reactions with alkyl halides in dimethylformamide (Helvenston and Castro 1992). The three classes of reaction were the same as those observed with Fe(II) deuteroporphyrin IX that have already been noted. Experiments that were carried out to compare nanoscale catalysts composed of Fe, Ni, and Co complexes of several porphyrins, or cyanocobalamin have already been noted (Dror et al. 2005).

THIOL REDUCTANTS It has been shown that thiols can carry out dehalogenation both in the presence and absence of quinone mediators: a. The formation of tetrachloroethene from hexachloroethane in the presence of 5-hydroxynaphtho-1,4-quinone (juglone) (Perlinger et al. 1996).

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b. The reduction of hexachloroethane by both HS− and polysulfides (S42−) to tetrachloroethene and pentachloroethane (Miller et al. 1998). c. The transformation of tetrachloromethane by HS− and cysteine in the presence of quinones as electron-transfer mediators has been examined (Doong and Chiang 2005). 5-Hydroxy-naphtho-1,4-quinone (juglone) was the most effective and much greater than anthraquinone 2,6-disulfate. The enhancement was the result of the formation of mercaptoquinones or, from anthraquinone 2,6-disulfate semiquinone radicals.

REDUCTIVE DESULFURIZATION In the context of diagenesis in recent anoxic sediments, reduced carotenoids, steroids, and hopanoids have been identified, and it has been suggested that reduction by sulfide, produced for example, by the reduction of sulfate could play an important part (Hebting et al. 2006). The partial reduction of carotenoids by sulfide has been observed as a result of the addition of sulfide to selected allylic double bonds, followed by reductive desulfurization. This is supported by the finding that the thiol in allylic thiols could be reductively removed by sulfide to produce unsaturated products from freeradical reactions (Hebting et al. 2003).

REDUCTIONS OTHER THAN DEHALOGENATION Aromatic nitro compounds include both important explosives and a number of agrochemicals. Concern with their fate has motivated extensive examination of their reduction to amines under a range of conditions. 1. Reduction of monocyclic aromatic nitro compounds has been demonstrated: (a) with reduced sulfur compounds mediated by a naphthoquinone or an iron porphyrin (Schwarzenbach et al. 1990), and (b) by Fe(II) and magnetite produced by the action of the anaerobic bacterium Geobacter metallireducens (Heijman et al. 1993). Quinone-mediated reduction of monocyclic aromatic nitro compounds by the supernatant monocyclic aromatic nitro compounds has been noted (Glaus et al. 1992), and these reactions may be significant in determining the fate of aromatic nitro compounds in reducing environments (Dunnivant et al. 1992). 2. The reduction of 2,4,6-trinitrotoluene with Fe0 has been extensively studied (references in Bandstra et al. 2005), and it has finally produced 2,4,6-triaminotoluene that could undergo polymerization. 3. The degradation of trifluralin [2,6-dinitro-N,N-dipropyl-4-(trifluoromethyl)benzenamine] by Fe(II)/goethite has been examined under anaerobic conditions as a model for reactions in flooded soils (Klupinski and Chin 2003). A range of transformation products was found including those produced by reduction of the nitro groups, dealkylation, and formation of benziminazoles presumably via intermediate nitroso compounds (Figure 1.26). A number of other abiotic reductions have been described: 1. The aerobic biodegradation of monocyclic azaarenes frequently involves reduction (Chapter 10, Part 1), but purely chemical reduction may take place under highly anaerobic conditions. This has been encountered with the substituted 1,2,4-triazolo[1,5a]pyrimidine (Flumetsulam) (Wolt et al. 1992) (Figure 1.27). 2. Cell-free supernatants may mediate reductions. The reduction of aromatic nitro compounds by SH− was mediated by the filtrate from a strain of Streptomyces sp. that is known to synthesize 2-amino-3-carboxy-5-hydroxybenzo-1,4-quinone (cinnaquinone)

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Abiotic Reactions

29 NO2 N

F3 C

NO2

NH2

N NO2 F3 C

NO2

N N

F3 C

FIGURE 1.26

Degradation of 2,6-dinitro-N,N-dipropyl(trifluoromethyl)benzenamine.

F

F N

N

N

NH−SO 2 N

F

FIGURE 1.27

N

NH−SO 2 N

CH 3

F

N

N H

CH 3

Reductive degradation of 1,2,4-triazolo[1,5a]pyrimidine.

and the 6,6′-diquinone (dicinnaquinone) as secondary metabolites (Glaus et al. 1992). The quinones presumably function as electron-transfer mediators. 3. Dimerization of a number of arylalkenes catalyzed by vitamin B12 and Ti(III) as reductant has been examined (Shey et al. 2002). Mechanisms were examined including the requirement of a reductant, and a reaction was proposed that involved the formation of radicals at the benzylic carbon atoms.

NUCLEOPHILIC REACTIONS Although the foregoing reactions involve dehalogenation by reduction or elimination, nucleophilic displacement of chloride may also be important. This has been examined with dihalomethanes using HS− at concentrations that might be encountered in environments where active anaerobic sulfate reduction is taking place. The rates of reaction with HS− exceeded those for hydrolysis and at pH values above 7 in systems that are in equilibrium with elementary sulfur, the rates with polysulfide exceeded those with HS−. The principal product from dihalomethanes was the polythiomethylene HS−(CH2–S)nH (Roberts et al. 1992). Attention is briefly drawn to hydrolytic procedures that have been considered for the destruction of xenobiotics. Although these are carried out under conditions that are not relevant to the aquatic environment, they may be useful as a background to alternative remediation programs. Three examples involving CaO and related compounds may be used as examples of important and unprecedented reactions: 1. The destruction of DDT by ball milling with CaO resulted in substantial loss of chloride and produced a graphitic product containing some residual chlorine. In addition, an exceptional rearrangement occurred with the formation of bis(4-chlorophenyl)ethyne that was identified by 1H NMR (Hall et al. 1996) (Figure 1.28). 2. Treatment of 1,2,3,4-tetrachlorodibenzo[1,4]dioxin on Ca-based sorbents at temperatures of 160–300°C resulted in its conversion into products with molecular masses of

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Cl

Cl

C

Cl

C

C

Cl

CCl2

FIGURE 1.28 Alkaline destruction of DDT. CHO CO 2 H

CHO NH 2

N

CO N CONH 2

N CONH 2

N

N CHO

O CHO

FIGURE 1.29 Ozonization of carbamazepine.

302 and 394 that were tentatively identified as chlorinated benzofurans and 1-phenylnaphthalene or anthracenes (Gullett et al. 1997). 3. CaO was activated by the addition of Ni (acetylacetonate)2 followed by calcining at 973 K. This resulted in the production of nickel on the carbonate matrix, and was effective in decomposing dichlorodifluoromethane (CFC-12) to CO2 at 723 K—with low amounts of CO—while the halogen was combined as CaFCl (Tamai et al. 2006).

OXIDATIONS These have already been noted in the context of hydroxyl radical-initiated oxidations, and reference should be made to an extensive review by Worobey (1989) that covers a wider range of abiotic oxidations. Some have attracted interest in the context of the destruction of xenobiotics, and reference has already been made to photochemically induced oxidations. Anodic Oxidation An interesting study examined the anodic oxidation of EDTA at alkaline pH on a smooth platinum electrode (Pakalapati et al. 1996). Degradation was initiated by removal of the acetate side chains as formaldehyde, followed by deamination of the ethylenediamine that formed glyoxal and oxalate. Oxalate and formaldehyde are oxidized to CO2, and adsoption was an integral part of the oxidation. Ozone Although reactions carried out by ozone have attracted enormous attention in the atmospheric environment, ozone has also been used extensively in the treatment of drinking water without the production of undesirable trihalomethanes from the use of molecular chlorine (Richardson et al. 1999). It has been examined for the removal of a number of contaminants, and ozone is considered to be a selective oxidant, even though quite complex reactions may occur. 1. Laboratory experiments on the ozonization of the dibenzoazepine drug carbamazepine (McDowell et al. 2005) showed the occurrence of a number of transformation products that were initiated by reaction at the olefinic double bond between the rings (Figure 1.29). 2. The degradation of aminodinitrotoluenes has been examined, and destruction of the rings took place with the release of NO3− and NO2−. Use of 15N amino compounds revealed that pyruvamide was formed from the 2-amino compound and oxamic acid from the 4-amino compound (Spanggord et al. 2000).

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3. Ozone attacks the rings of PAHs rather indiscriminately with fission of the rings to produce aldehyde groups. There has been concern, however, since the products may be more harmful than their precursors. In the studies that are used as illustration, in vitro gap junctional intracellular communication (GJIC) was used to assess adverse alteration on the expression of genes at the transcription, translational, or posttranslational level: a. Pyrene was degraded to two groups of compounds: phenanthrene rings with one or more carbonyl groups by oxidation and fission of ∆9–10, and biphenyls with four carbonyl groups ortho to the ring junction by fission of both ∆9–10 and ∆4–5 (Herner et al. 2001). b. The product from chrysene was produced by oxidation and fission of ∆11–12 to the dialdehyde (Luster-Teasley et al. 2002). 4. Effective removal of the estrogens 17β-estradiol, estrone, and 17α-ethynylestradiol has been achieved using ozone under conditions that simulated those used for water treatment (Deborde et al. 2005). Analysis of the products showed the occurrence of two reactions: a. Fission of rings A and B in estrone and 17β-estradiol to yield products with a carboxyl group at C9 and a carboxyethyl group at C8 (Figure 1.30a). b. Oxidation of ring A to a 1,2-dione that underwent a benzylic acid rearrangement with ring contraction to a cyclopentane carboxylate (Figure 1.30b). Analogous reactions were postulated to occur for 17α-ethynylestradiol and, on the basis of experiments with model compounds, oxidation of the ethynyl group to –COCHO (Huber et al. 2004). 5. Combined treatment of atrazine with ozone and H2O2 resulted in retention of the triazine ring, and oxidative dealkylation with or without replacement of the 2-chloro group by hydroxyl (Nélieu et al. 2000). Reaction with ozone and hydroxyl radicals formed the analogous products with the additional formation of the acetamido group from one of the N-alkylated groups (Acero et al. 2000). 6. A range of products was obtained from aniline including those from oxidative coupling (azozybenzene, azobenzene, and benzidine), and phenazine by dimerization (Chan and Larson 1991). Oxidation of m-phenylenediamine was initiated by the oxidation of two molecules to produce an N-phenyl-2-aminoquinone-imine that reacted with m-phenylenediamine to produce 2-amino-5-phenylaminoquinone-imine after further oxidation (Kami et al. 2000).

OH

HO 2 C OH (a)

CO2 H OH

OH

(b) HO O

HO 2C HO

O

FIGURE 1.30

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(a) and (b) Ozonization of estrone/estradiol.

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Hydrogen Peroxide The use of hydrogen peroxide in conjunction with Fe(II) (Fenton’s reagent) or ozone has already been noted. It has been used alone to examine the products from o- and m-phenylenediamines in the context of their mutagenicity (Watanabe et al. 1989). Successive reactions produced 3,4-diaminophenazine from o-phenylenediamine, and 3,7-diaminophenazine from m-phenylenediamine. Manganese Dioxide Manganese dioxide has been used to carry out a range of chemical oxidations, and is the reagent of choice for the oxidation of allylic alcohols (Hudlický 1990). There are several methods for its preparation that may account for differences in its activity. Although it has seldom been exploited for the oxidation of contaminants, two widely different applications have been described: 1. The transformation of several fluoroquinolone antibacterials was examined and a number of products from ciprafloxacin were tentatively identified (Zhang and Huang 2005). The quinolone ring was unchanged, and the major product was produced by fission of the piperazine ring to an aniline. 2. The oxidation of triclosan was initiated by oxidation to the PhO• radical. In analogy with reactions established in biomimetic synthesis, this underwent coupling to produce biphenyl ethers and biphenyls that were oxidized further to diphenoquinones (Zhang and Huang 2003). Sulfate Radicals The photochemical generation of sulfate radicals for the degradation of perfluorocarboxylic acids has already been noted (Hori et al. 2005a,b). Sulfate radicals are powerful oxidants, and can be produced by the Co-mediated degradation of peroxymonosulfate (HSO5−). They have been shown to be effective for the degradation of 2,4-dichlorophenol and atrazine with the formation of sulfate (Anipsitakis and Dionysiou 2003), and it was shown that they were more effective than conventional Fenton oxidation with hydroxyl radicals, and were effective over a wider range of pH. Sulfate radicals are powerful oxidizing agents and can oxidize chloride to chlorine radicals and molecular chlorine that may be incorporated into the reactant. For example, reaction of peroxymonosulfate with 2,4-dichlorophenol in the presence of chloride produced 2,4,6-trichlorophenol; 2,3,5,6-tetrachlorohydroquinone; and the ring fission products penta- and tetrachloroacetone, and tetrachloromethane (Anipsitakis et al. 2006). Dioxirane that is produced in situ by the reaction of methylpyruvate and peroxymonosulfate has been evaluated as (a) an effective disinfecting agent (Wong et al. 2006) and (b) oxidant of alkenes, alkanes, and phenols (references in Wong et al. 2006).

HALOGENATION Halogenation is important for disinfecting drinking water supplies, generally using molecular chlorine. Most attention has been directed to the adverse production of haloforms and haloacetates from reactions of chlorine with natural substrates, although in water containing bromide/iodide, a number of other reactions may occur. a. While attempts to decrease the formation of haloforms have used chloramination, the undesirable formation of iodoform may take place in the presence of iodide (Leitner 1998). b. The formation of iodoacetate and iodinated propenoates has been demonstrated (Plewa et al. 2004), and the cytotoxic and genotoxic properties of the former (Cemeli et al. 2006) has aroused concern.

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c. In water containing particularly high concentrations of bromide, there is a rather special situation; chlorination with a mixture of chlorine and ClO2 produced 2,3,5tribromopyrrole—putatively from humic acids—that is strongly cytotoxic and genotoxic (Richardson et al. 2003). In addition, there is interest in the halogenation of a wide range of anthropogenic contaminants, some of which may occur in raw water before treatment. 1. Pyrene is a common PAH contaminant and may occur in drinking water. Chlorination of water with or without bromide that may be present in coastal environments has been examined. Both chlorinated and brominated pyrenes with halogen substituents at the 1,3-, 1,6-, and 1,8-positions were found, and could putatively be produced by reaction of pyrene with hypochlorous acid and hypochlorite (Hu et al. 2006). 2. Degradation of the insecticide chlorpyrifos has been examined in solutions of aqueous chlorine in which the primary oxidant is hypochlorous acid/hypochlorite. The final product was 2,3,5-trichloropyrid-2-one produced either directly, or via initial oxidative conversion of the thioate ester by replacement of the sulfur with oxygen (Duirk and Collette 2006). An analogous oxidation was found with diazinon (Zhang and Pehkonen 1999). 3. Transformation of the widely used over-the-counter analgesic acetaminophen (paracetamol) during chlorination produced the toxic 1,4-benzoquinone via the N-acetylquinone-imine and minor amounts of products from chlorination of the phenolic ring (Bedner and Maccrehan 2006). 4. The chlorination of the antibacterial sulfonamide sulfamethoxazole was initiated by Nchlorination of the primary amine. Further reaction of the N,N-dichlorinated compound resulted in the final production of 3-amino-5-methyloxazole and 1,4-benzoquinone-imine (Dodd and Huang 2004).

Cl (a)

A

A

HO

A

+

HO

HO Cl

Cl

O (b)

C

O

D

C

CO 2 H Cl

D

C

Cl C (c)

Cl

A HO

A O

C

Cl A

B HO

O C

FIGURE 1.31

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Chlorination of estrone.

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5. The effect of chlorinating estrone in aqueous solution has been examined. Three reactions involving chlorination of different rings emerged: i. Ring A to produce 2- and 4-chlorinated products (Figure 1.31a) ii. Ring D at C16 followed by hydrolytic ring fission of ring D (Figure 1.31b) iii. Ring A at C10 followed by fission of ring B with the formation of a ketone in ring C at C9 (Figure 1.31c) (Hu et al. 2003). 6. Amino acids have been found in samples of river water (Lee et al. 2006), and chlorination of glycine may produce cyanogen chloride, via N,N-dichloroglycine. At pH values >6, this is converted into cyanogen chloride, whereas at lower pH values N-chloromethylimine is formed (Na and Olson 2006).

THERMAL REACTIONS Gas Phase The products of incomplete combustion may be associated with particulate matter before their discharge into the atmosphere, and these may ultimately enter the aquatic and terrestrial environments in the form of precipitation and dry deposition. It is therefore essential to ensure total destruction of the contaminants, generally by raising the temperature. The spectrum of compounds that have been examined is quite extensive, and several of them are produced by reactions between hydrocarbons and inorganic sulfur or nitrogen constituents of air. Some illustrative examples involving other types of reaction include the following: 1. The pyrolysis of vinylidene chloride produced a range of chlorinated aromatic compounds including polychlorinated benzenes, styrenes, and naphthalenes (Yasahura and Morita 1988), and a series of chlorinated acids including chlorobenzoic acids has been identified in emissions from a municipal incinerator (Mowrer and Nordin 1987). 2. Nitroaromatic compounds have been identified in diesel engine emissions (Salmeen et al. 1984), and attention has been directed particularly to 1,8- and 1,6-dinitropyrene that are mutagenic, and possibly carcinogenic (Nakagawa et al. 1983). 3. A wide range of azaarenes including acridines and benzacridines, 4-azafluorene, and 10-azabenzo[a]pyrene (Figure 1.32) has been identified in particulate samples of urban air, and some of them have been recovered from contaminated sediments (Yamauchi and Handa 1987). 4. Ketonic and quinonoid derivatives of aromatic hydrocarbons have been identified in automobile (Alsberg et al. 1985) and diesel exhaust particulates (Levsen 1988), and have been recovered from samples of marine sediments (Fernández et al. 1992). 5. Halogenated phenols, particularly 2-bromo-, 2,4-dibromo-, and 2,4,6-tribromophenol, have been identified in automotive emissions and are the products of thermal reactions involving the dibromoethane fuel additive (Müller and Buser 1986). It could therefore no longer be assumed that such compounds are exclusively the products of biosynthesis by marine algae. 6. Complex reactions occur during high-temperature treatment of aromatic hydrocarbons. An important class of reactions involves the cyclization and condensation of simpler

N

FIGURE 1.32

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N

Azaarenes identified in particulate samples of urban air.

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CH 3

CH 3

CH 3

FIGURE 1.33

Products from the pyrolysis of 2-methylnaphthalene.

PAHs to form highly condensed polycyclic compounds. This is discussed more fully by Zander (1995). a. A number of pentacyclic aromatic hydrocarbons have been identified as products of the gas-phase pyrolysis of methylnaphthalenes. These were formed from 1-methyl- and 2-methylnaphthalene by dimerization (Lang and Buffleb 1958) at various positions, whereas direct coupling with loss of the methyl group was found to be dominant with 2-methylnaphthalene (Lijinsky and Taha 1961) (Figure 1.33). b. A hypothetical scheme involving 2-carbon and 4-carbon additions has been used to illustrate the formation of coronene (circumbenzene) and ovalene (circumnaphthalene) from phenanthrene (Figure 1.34). 7. Concern has been expressed over the formation of chlorinated dibenzo[1,4]dioxins and dibenzofurans during the thermal transformation of organic material in the presence of chloride or organochlorine compounds. a. The high-temperature oxidation of 2-chlorophenol yielded a number of products. At 600°C these included 4,6-dichlorodibenzofuran, 1-chlorodibenzo[1,4]dioxin, and low concentrations of 2,4- and 2,6-dichlorophenol that were produced by reactions with chloride radicals and hydroxyl radicals. They were, however, essentially eliminated at temperatures above 800°C (Evans and Dellinger 2005a). In contrast, under comparable conditions, although a mixture of 2-chloro- and 2-bromophenol also yielded a range of halogenated dibenzofurans, 4,6-dibromodibenzofuran was not produced (Evans and Dellinger 2005b). b. In experiments to clarify the ameliorating effect on the formation of these undesirable products, the effect of adding urea was studied. Reaction of graphite, CuCl2, and urea at 300°C resulted in substantial reduction in the amounts of chlorinated dioxins and furans that were produced, and the formation of 2,6-dichlorobenzonitrile, 4-chloro2-methylaniline, and pentachloropyridine (Kuzuhara et al. 2005).

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C2

C4

C2

C2

C2

C2

FIGURE 1.34

C2

Successive C2 and C4 addition reactions.

In Water There are two distinct conditions that have been used: above the critical temperature and pressure (374°C and 218 atm) water becomes a supercritical fluid in which the distinction between the liquid and gaseous states disappears. Since supercritical water can dissolve nonpolar compounds, it has been examined for the degradation of such contaminants. Subcritical water in which the liquid state is maintained by the pressure of the containing vessel has also achieved attention. 1. Destruction of the explosives RDX, HMX, and TNT has been examined using subcritical water in both laboratory- and pilot-scale experiments. In contaminated soils at 150°C, considerable amounts of TNT remained in the soil after reaction for 5 h, and of HMX for 2.5 h. In the pilot-scale experiments, heating at 275°C for 1 h accomplished complete destruction of RDX and TNT, and ca. 98% destruction of HMX (Hawthorne et al. 2000).

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2. PCBs a. Destruction of PCBs containing congeners with two to eight chlorine substituents was examined in supercritical water under oxidizing or alkaline reducing conditions. The latter were more effective and, although chlorinated dibenzofurans with one to six chlorine substituents were formed at temperatures of 250°C they were removed at 350°C. Experiments with the 245-245 hexachloro congener were used to illustrate the formation of chlorinated dibenzofurans (Weber et al. 2002). b. Examination of the destruction of Arochlor 1254 in paint scrapings at 350°C revealed the significance of the metals. Experiments with zero-valent metal additions in the absence of paint showed their effectiveness in the order Pb≡Cu >Al >Zn >Fe. With Pb and Cu destruction of tetra to heptachloro congeners to predominantly monochlorobiphenyl occurred (Kubátova et al. 2003). c. Zero-valent iron has been used for the reductive dechlorination of PCBs in subcritical water at 250°C (Yak et al. 1999). Extensive loss of the more highly chlorinated congeners in Arochlor 12609 took place with the formation of congeners having one to four chlorine substituents, while with longer exposure times these were virtually eliminated. 3. At 200°C destruction of γ-hexachlorocyclohexane in subcritical water occurred in 1 h, and for dieldrin in 1 h at 300°C. The pathway for γ-HCH involved successive formation of 1,2,4-trichloro-, 1,4-dichloro-, and monochlorobenzene that was hydrolyzed to phenol (Kubátova et al. 2002). 4. The use of subcritical water at 275°C was successful in removing PAHs with two to six rings (including the carcinogenic benzo[a]pyrene) from soil at a contaminated site, and the wastewater that was produced appeared to be suitable for further exploitation (Lagadec et al. 2000).

ELECTROCATALYTIC REACTIONS The energy for the fission of the covalent bond in organic contaminants is normally supplied thermally using thermodynamically accessible chemical or biochemical reactions, or by the introduction of catalysts to lower the activation energy of the reactions. There has been interest, however, in using electrical energy in a number of forms to carry out these reactions. A selection of processes for the destruction of contaminant is noted with some illustrative examples. 1. The simplest applications involve direct anodic oxidation in aqueous media. a. During the anodic oxidation of EDTA at alkaline pH on a smooth platinum electrode (Pakalapati et al. 1996), degradation was initiated by removal of the acetate side chains as formaldehyde, followed by deamination of the ethylenediamine that is formed to glyoxal and oxalate. Oxalate and formaldehyde were oxidized to CO2. b. Using a platinum electrode coated with a metal oxide film (Ti, Ru, Sn, and Sb) in conjunction with a solid Nafion membrane, complete destruction of 4-chlorophenol could be achieved in reactions involving the formation of benzoquinone followed by ring fission to succinate acetate and CO2 (Johnson et al. 1999). On account of the products this was termed “electrochemical incineration.” 2. Electrohydraulic discharge plasmas have been used in different configurations. a. The degradation of 2,4,6-trinitrotoluene has been carried out in the presence of ozone. The hydroxyl radicals that were produced brought about oxidation to an intermediate that was identified as 2,4,6-trinitrobenzaldehyde (Lang et al. 1998). b. The degradation of 4-chlorophenol and 3,4-dichloroaniline was carried out by the shock wave that was developed in the cell (Willberg et al. 1996).

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3. Variants of glow discharge plasmas have been used: a. The dechlorination of pentachlorophenol took place with the formation of C1 and C2 carboxylic acids (Sharma et al. 2000). b. In the presence of Fe(III), degradation of phenol which took place by hydroxylation followed by ring fission (Liu and Jiang 2005) in reactions are analogous to those in the electro-Fenton reaction that has been examined for the degradation of 4-nitrophenol (Oturan et al. 2000). 4. Electrolytic reduction has been carried out under several conditions. a. Using a porous nickel electrode and at voltages that were maintained by a potentiostat, a number of reductive reactions were observed: tetrachloroethene and trichloroethene → ethane and ethene, and penta-, tetra-, tri, and dichloroethanes to ethane (Liu et al. 2000). b. Fuel cells have been used in various modifications to carry out gas-phase reductive dechlorination of trichloroethene: • Using a polymer electrolyte membrane cell in which H2 flowed through the anode chamber. The major intermediate chlorinated products from tetrachloroethene or tetrachloromethane were trichloroethene or trichloromethane, and these were finally reduced to a mixture of ethane and ethene, or methane (Liu et al. 2001). • Reduction of trichloroethene to ethane took place in a modified fuel cell to which H2 was introduced, although the loss of catalytic activity with time could present a serious limitation (Ju et al. 2006).

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Atkinson R, J Arey, B Zielinska, SM Aschmann (1987b) Kinetics and products of the gas-phase reactions of OH radicals and N2O5 with naphthalene and biphenyl. Environ Sci Technol 21: 1014–1022. Atkinson R, SM Aschmann, AM Winer, JN Pitts (1985a) Kinetics and atmospheric implications of the gasphase reactions of NO3 radicals with a series of monoterpenes and related organics at 294 ± 2 K. Environ Sci Technol 19: 159–163. Atkinson R, SM Aschmann, AM Winer, WPL Carter (1985b) Rate constants for the gas-phase reactions of NO3 radicals with furan, thiophene, and pyrrole at 295 ± 1 K and atmospheric pressure. Environ Sci Technol 19: 87–90. Atkinson R, SM Aschmann, AW Winer (1987c) Kinetics of the reactions of NO3 radicals with a series of aromatic compounds. Environ Sci Technol 21: 1123–1126. Baker JI, RA Hites (2000) Is combustion the major source of polychlorinated dibenzo-p- dioxins and dibenzofurans to the environment? A mass balance study. Environ Sci Technol 34: 2879–2886. Bandstra JZ, R Miehr, RL Johnson, PG Tratnyek (2005) Reduction of 2,4,6-trinitrotoluene by iron metal: kinetic controls on product distributions in batch experiments. Environ Sci Technol 39: 230–238. Bank S, RJ Tyrrrell (1984) Kinetics and mechanism of alkaline and acidic hydrolysis of aldicarb. J Agric Food Chem 32: 1223–1232. Barbas JT, ME Sigman, R Dabestani (1996) Photochemical oxidation of phenanthrene sorbed on silica gel. Environ Sci Technol 30: 1776–1780. Baretto RD, KA Gray, K Anders (1995) Photocatalytic degradation of methyl-tert-butyl ether in TiO2 slurries: a proposed reaction scheme Water Res 29: 1243–1248. Bedner M, WA Maccrehan (2006) Transformation of acetaminophen by chlorination produces the toxicants 1,4-benzoquinone and N-acetyl-p-benzoquinone imine. Environ Sci Technol 40: 516–522. Beitz T, W Bechmann, R Mitzner (1998) Investigations of reactions of selected azaarenes with radicals in water. 1. Hydroxyl and sulfate radicals. J Phys Chem A 102: 6760–6765. Belly RT, JJ Lauff, CT Goodhue (1975) Degradation of ethylenediaminetetraacetic acid by microbial populations from an aerated lagoon. Appl Microbiol 29: 787–794. Beltran E, H Fenet, JF Cooper, CM Coste (2000) Kinetics of abiotic hydrolysis of isoxaflutole: influence of pH and temperature in aqueous mineral buffered medium. J Agric Food Chem 48: 4399–4403. Bergon M, NB Hamida, J-P Calmon (1985) Isocyanate formation in the decomposition of phenmedipham in aqueous media. J Agric Food Chem 33: 577–583. Bláha L, J Klánová, P Klán, J Janosek, M Skarek, R Rúzicka (2004) Toxicity increases in ice containing monochlorophenols upon photolysis: environmental consequences. Environ Sci Technol 38: 2873–2878. Bosma TNP, FHM Cottaar, MS Posthumus, CJ Teunis, A van Veidhuizen, G Schraa, AJB Zehnder (1994) Comparison of reductive dechlorination of hexachloro-1,3-butadiene in Rhine sediments and model systems with hydroxocobalamin. Environ Sci Technol 28: 1124–1128. Bosma TNP, JR van der Meer, G Schraa, ME Tros, AJB Zehnder (1988) Reductive dechlorination of all trichloro- and dichlorobenzene isomers. FEMS Microbiol Ecol 53: 223–239. Brezonik PL, J Fulkerson-Brekken (1998) Nitrate-induced photoysis in natural waters: controls on concentrations of hydroxyl radical photo-intermediates by natural scavenging agents. Environ Sci Technol 32: 3004–3010. Brownlee BG, JH Carey, GA MacInnes, IT Pellizzari (1992) Aquatic environmental chemistry of 2-(thiocyanomethylthio)benzothiazole and related benzothiazoles. Environ Toxicol Chem 11: 1153–1168. Brubaker WW, RA Hites (1997) Polychlorinated dibenzo-p-dioxins and dibenzofurans: gas-phase hydroxyl radical reactions and related atmospheric removal. Environ Sci Technol 31: 1805–1810. Brubaker WW, RA Hites (1998) Gas-phase oxidation products of biphenyl and polychlorinated biphenyls. Environ Sci Technol 32: 3913–3918. Bunce NJ, L Liu, J Zhu, DA Lane (1997) Reaction of naphthalene and its derivatives with hydroxyl radicals in the gas phase. Environ Sci Technol 31: 2252–2259. Burhenne J, M Ludwig, M Spiteller (1997a) Photolytic degradation of fluoroquinolone carboxylic acids in aqueous solution. Primary photoproducts and half-lives. Environ Sci Pollut Res 4: 10–15. Burhenne J, M Ludwig, M Spiteller (1997b) Photolytic degradation of fluoroquinolone carboxylic acids in aqueous solution. Isolation and structural elucidation of polar photometabolites. Environ Sci Pollut Res 4: 61–67. Burns SE, JP Hassett, MV Rossi (1996) Binding effects on humic-mediated photoreaction: intrahumic dechlorination of mirex in water. Environ Sci Technol 30: 2934–2941. Burns SE, JP Hassett, MV Rossi (1997) Mechanistic implications of the intrahumic dechlorination of mirex. Environ Sci Technol 31: 1365–1371.

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Moore DE, S Roberts-Thomson, D Zhen, CC Duke (1990) Photochemical studies on the anti-inflammatory drug diclofenac. Photochem Photobiol 52: 685–690. Mowrer J, J Nordin (1987) Characterization of halogenated organic acids in flue gases from municipal waste incinerators. Chemosphere 16: 1181–1192. Müller MD, H-R Buser (1986) Halogenated aromatic compounds in automotive emissions from leaded gasoline additives. Environ Sci Technol 20: 1151–1157. Müller SR, H-R Zweifel, DJ Kinnison, JA Jacobsen, MA Meier, MM Ulrich, RP Schwarzenbach (1996) Occurrence, sources, and fate of trichloroacetic acid in Swiss waters. Environ Toxicol Chem 15: 1470–1478. Na C, TM Olson (2006) Mechanism and kinetics of cyanogen chloride formation from the chlorination of glycine. Environ Sci Technol 40: 1469–1477. Nakagawa R, S Kitamori, K Horikawa, K Nakashima, H Tokiwa (1983) Identification of dinitropyrenes in diesel-exhaust particles. Their probable presence as the major mutagens. Mutation Res 124: 201–211. Nélieu S, L Kerhoas, J Einhorn (2000) Degradation of atrazine into ammeline by combined ozone/hydrogen peroxide treatment in water. Environ Sci Technol 34: 430–437. Nelson L, I Shanahan, HW Sidebottom, J Treacy, OJ Nielsen (1990) Kinetics and mechanism for the oxidation of 1,1,1-trichloroethane. Int J Chem Kinet 22: 577–590. Nishioka MG, CC Howard, DA Conros, LM Ball (1988) Detection of hydroxylated nitro aromatic and hydroxylated nitro polycyclic aromatic compounds in ambient air particulate extract using bioassay-directed fractionation. Environ Sci Technol 22: 908–915. Nomiyama K, T Tanizaki, H Ishibashi, K Arizono, R Shinohara (2005) Production mechanism of hydroxylated PCBs by oxidative degradation of selected PCBs using TiO2 in water and estrogenic activity of their intermediates. Environ Sci Technol 39: 8762–8769. Nörtemann B (1992) Total degradation of EDTA by mixed cultures and a bacterial isolate. Appl Environ Microbiol 58: 671–676. Orlando JJ (2003) Atmospheric chemistry of organic bromine and iodine compounds. Handbook Environ Chem 3R: 253–299. Oturan MA, J Peiroten, P Chartin, AJ Acher (2000) Complete destruction of p-nitrophenol in aqueous medium by electro-Fenton method. Environ Sci Technol 34: 3474–3479. Paciolla MD, G Davies, SA Jansen (1999) Generation of hydroxyl radicals from metal-loaded humic acids. Environ Sci Technol 33: 1814–1818. Pakalapati SNR, BN Popov, RE White (1996) Anodic oxidation of ethylenediaminetetraacetic acid on platinum electrode in alkaline medium. J Electrochem Soc 143: 1636–1643. Pelizzetti E, V Maurino, C Minero, V Carlin, E Pramauro, O Zerbinati, ML Tosata (1990) Photocatalytics degradation of atrazine and other s-triazine herbicides. Environ Sci Technol 24: 1559–1565. Perdue EM, NL Wolfe (1982) Modification of pollutant hydrolysis kinetics in the presence of humic substances. Environ Sci Technol 16: 847–852. Perlinger JA, W Angst, RP Schwarzenbach (1996) Kinetics of the reduction of hexachloroethane by juglone in solutions containing hydrogen sulfide. Environ Sci Technol 30: 3408–3417. Pitts JN Jr, B Zielinska, JA Sweetman, R Atkinson, AM Winer (1985b) Reactions of adsorbed pyrene and perylene with gaseous N2O5 under simulated atmospheric conditions. Atmos Environ 19: 911–915. Pitts N Jr, R Atkinson, JA Sweetman, B Zielinska (1985a) The gas-phase reaction of naphthalenes with N2O5 to form nitronaphthalenes. Atmos Environ 19: 701–705. Platt UF, AM Winer, HW Biermann, R Atkinson, JN Pitts, Jr (1984) Measurement of nitrate radical concentrations in continental air. Environ Sci Technol 18: 365–369. Platz J, OJ Nielsen, J Sehested, TJ Wallington (1995) Atmospheric chemistry of 1,1,1-trichloroethane: UV absorption spectra and self-reaction kinetics of CCl3CH2 and CCl3CH2O2 radicals, kinetics of the reactions of the CCl3CH2O2 radical with NO and NO2, and the fate of alkoxy radical CCl3CH2O. J Phys Chem 99: 6570–6579. Plewa MJ, ED Wagner, SD Richardson, AD Thurston, Y-T Woo, AB McKague (2004) Chemical and biological characterization of newly discovered iodoacid drinking water disinfection byproducts. Environ Sci Technol 38: 4713–4722. Pradhan SP, JR Paterek, BY Liu, JR Conrad, VJ Srivastava (1997) Pilot-scale bioremediation of PAH-contaminated soils. Appl Biochem Biotechnol 63/65: 759–773. Quinn J, C Geiger, C Clausen, K Brooks, C Coon, S O´Hara, T Krug, D Major, W-S Yoon, A Gavskar, T Holdsworth (2005) Field demonstration of DNAPL dehalogenation using emulsified zero-valent iron. Environ Sci Technol 39: 1309–1318.

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Zepp RG, GL Baugham, PA Scholtzhauer (1981a) Comparison of photochemical behaviour of various humic substances in water. I. Sunlight induced reactions of aquatic pollutants photosensitized by humic substances. Chemosphere 10: 109–117. Zepp RG, GL Baugham, PA Scholtzhauer (1981b) Comparison of photochemical behaviour of various humic substances in water: II Photosensitized oxygenations. Chemosphere 10: 119–126. Zepp RG, PF Schlotzhauer (1983) Influence of algae on photolysis rates of chemicals in water. Environ Sci Technol 17: 462–468. Zepp RG, PF Schlotzhauer, RM Sink (1985) Photosensitized transformations involving energy transfer in natural waters: role of humic substances. Environ Sci Technol 19: 74–81. Zhang H, C-H Huang (2003) Oxidative transformation of triclosan and chlorophene by manganese dioxides. Environ Sci Technol 37: 2421–2430. Zhang H, C-H Huang (2005) Oxidative transformation of fluoroquinolone antibacterial agents and structuiralöly related amines by manganese dioxide. Environ Sci Technol 39: 4474–4483. Zhang Q, SO Pehkonen (1999) Oxidation of diazinon by aqueous chlorine: kinetics, mechanisms, and product studies. J Agric Food Chem 47: 1760–1766. Zielinska B, J Arey, R Atkinson, PA McElroy (1988) Nitration of acephenanthrylene under simulated atmospheric conditions and in solution and the presence of nitroacephenanthrylenes in ambient particles. Environ Sci Technol 22: 1044–1048. Zielinska B, J Arey, R Atkinson, PA McElroy (1989) Formation of methylnitronaphthalenes from the gasphase reactions of 1- and 2-methylnaphthalene with hydroxyl radicals and N2O5 and their occurrence in ambient air. Environ Sci Technol 23: 723–729. Zoro JA, JM Hunter, G Eglinton, CC Ware (1974) Degradation of p,p´-DDT in reducing environments. Nature 247: 235–237.

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Biotic Reactions

PART 1 MICROBIAL REACTIONS INTRODUCTION It is generally conceded that biotic reactions involving microorganisms are of major significance in determining the fate and persistence of organic compounds in aquatic and terrestrial ecosystems. The role of higher biota in carrying out important transformations is addressed in Part 2 of this chapter. Archaea, bacteria, cyanobacteria, fungi, yeasts, and algae comprise a large and diverse number of taxa. However, only a relatively small number of even the genera have been examined in the context of biodegradation, and there is no way of determining how representative these are. Care should therefore be exercised in drawing conclusions about the metabolic capability of the plethora of taxa included within these major groups of microorganisms. Attention is directed to the role of populations that are discussed in Chapter 13.

DEFINITIONS—DEGRADATION AND TRANSFORMATION It is essential to make a clear distinction between biodegradation and biotransformation in the beginning: Biodegradation. Under aerobic conditions, biodegradation results in the mineralization of an organic compound to carbon dioxide and water and—if the compound contains nitrogen, sulfur, phosphorus, or chlorine—with the release of ammonium (or nitrite), sulfate, phosphate, or chloride. These inorganic products may then enter well-established geochemical cycles. Under anaerobic conditions, methane may be formed in addition to carbon dioxide, and sulfate may be reduced to sulfide. Biotransformation. In contrast, biotransformation involves only a restricted number of metabolic reactions, and the basic framework of the molecule remains essentially intact. Illustrative examples of biotransformation reactions include the following, although it should be emphasized that other microorganisms may be able to degrade the substrates: 1. The hydroxylation of 14-chlorodehydroabietic acid by fungi (Figure 2.1) (Kutney et al. 1982) 2. The epoxidation of alkenes by bacteria (Patel et al. 1982; van Ginkel et al. 1987) 3. The formation of 16-chlorohexadecyl-16-chlorohexadecanoate from hexadecyl chloride by Micrococcus cerificans (Kolattukudy and Hankin 1968) CH3(CH2)14 –CH2Cl → ClCH2(CH2)14 –CH2–O–CO–(CH2)14 –CH2Cl 4. The O-methylation of chlorophenols to anisoles by fungi (Gee and Peel 1974; Cserjesi and Johnson 1972) and by bacteria (Suzuki 1978; Rott et al. 1979; Neilson et al. 1983; Häggblom et al. 1988) 5. The formation of glyceryl-2-nitrate from glyceryl trinitrate by Phanerochaete chrysosporium (Servent et al. 1991) 51

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HO Cl

Cl

CO 2 H

CO 2 H H

OH OH

HO

OH HO

Cl

CO 2 H

FIGURE 2.1

Cl

CO 2 H

Biotransformation of dehydroabietic acid by Mortierella isabellina.

The initial biotransformation products may, in some cases, be incorporated into cellular material. For example, the carboxylic acids formed by the oxidation of long-chain n-alkyl chlorides were incorporated into cellular fatty acids by strains of Mycobacterium sp. (Murphy and Perry 1983). The fungus Mortierella alpina, in which hexadecene was oxidized by the fungus by ω-oxidation (Shimizu et al. 1991), formed lipids that contained carboxylic acids containing both 18 and 20 carbon atoms including the unusual polyunsaturated acid 5cis, 8cis, 11cis, 14cis, 19-eicosapentaenoic acid. Metabolites of metolachlor that could only be extracted from the cells with acetone were apparently chemically bound to unidentified sulfur-containing cellular components (Liu et al. 1989). Biodegradation and biotransformations are of course alternatives, but they are not mutually exclusive. For example, it has been suggested that for chlorophenolic compounds, the O-methylation reaction may be an important alternative to reactions that bring about their degradation (Allard et al. 1987). Apart from the environmental significance of biotransformation reactions, many of them have enormous importance in biotechnology for the production of valuable metabolites, for example, the synthesis of hydroxylated steroids and in reactions that take advantage of the oxidative potential of methanotrophic bacteria (Lidstrom and Stirling 1990) and rhodococci (Finnerty 1992). It is also important to consider the degradation of xenobiotics in the wider context of metabolic reactions carried out by the cell. The cell must obtain energy to carry out essential biosynthetic (anabolic) reactions for its continued existence, and to enable growth and cell division to take place. The substrate therefore cannot be degraded entirely to carbon dioxide or methane, for example, and a portion must be channeled into the biosynthesis of essential molecules. Indeed, many organisms will degrade xenobiotics only in the presence of a suitable more readily degraded growth substrate that supplies both cell carbon and the energy for growth: this is relevant to discussions of “cometabolism” and “concurrent metabolism,” which is discussed in Chapter 4. Growth under anaerobic conditions is demanding both physiologically and biochemically, since the cells will generally obtain only low yields of energy from the growth substrate, and must, in addition, maintain a delicate balance between oxidative and reductive processes. Only a few examples of mechanisms for ATP generation in anaerobes are given below: 1. Clastic reactions from 2-keto acid-CoA esters produced in a number of degradations 2. Reactions involving carbamyl phosphate in the degradation of arginine in clostridia, and the fermentation of allantoin by Streptococcus allantoicus 3. The activity of formyl THF synthase during the fermentation of purines by clostridia 4. The reductive dechlorination of 3-chlorobenzoate by Desulfomonile tiedjei DCB-1

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5. The proton pump in Oxalobacter formigenes 6. The biotin-dependent carboxylases that couple the decarboxylation of malonate to acetate in Malonomonas rubra to the transport of Na+ across the cytoplasmic membrane True fermentation implies that a single substrate is able to provide carbon for cell growth and at the same time satisfy the energy requirements of the cell. A simple example of fermentation is the catabolism of glucose by facultatively anaerobic bacteria to pyruvate that is further transformed into a variety of products including acetate, butyrate, propionate, or ethanol by different organisms. In contrast, a range of electron acceptors may be used under anaerobic conditions to mediate oxidative degradation of the carbon substrate at the expense of the reduction of the electron acceptors. For example, the following reductions may be coupled to oxidative degradation: nitrate to nitrogen (or nitrous oxide), sulfate to sulfide, carbonate to methane, fumarate to succinate, trimethylamineN-oxide to trimethylamine, or dimethyl sulfoxide to dimethyl sulfide, which are discussed in Chapter 3, Part 2. The environments required by the relevant organisms are determined by the redox potential of the relevant reactions, so that increasingly reducing conditions are required for reduction of nitrate, sulfate, and carbonate. Attention is drawn to the dechlorination by anaerobic bacteria of both chlorinated ethenes and chlorophenolic compounds that serve as electron acceptors with electron donors including formate, pyruvate, and acetate. This is termed dehalorespiration and is important in the degradation of a range of halogenated compounds under anaerobic conditions, and is discussed further in Chapter 3, Part 2 and Chapter 7, Part 3. Probably, most of the microbial degradations and transformations that are discussed in this book are carried out by heterotrophic microorganisms that use the xenobiotic as a source of both carbon and energy, or by cometabolism (Chapter 4). In addition, xenobiotics may serve only as sources of nitrogen, sulfur, or phosphorus. The attention is briefly drawn here to those groups of organisms many of whose members are autotrophic or lithotrophic. Discussion of the complex issue of organic nutrition of chemolithotrophic bacteria, and the use of the term “autotrophy” is given in a classic review by Matin (1978). The groups of organisms that are discussed here in the context of biotransformation include 1. Ammonia-oxidizing bacteria including Nitrosomonas europaea 2. Facultatively heterotrophic thiobacilli that use a number of organic sulfur compounds as energy source 3. Oxygenic photolithotrophic algae and cyanobacteria It is important to underscore the fact that carbon dioxide is required not only for the growth of strictly phototrophic and lithotrophic organisms. Many heterotrophic organisms that are heterotrophic have an obligate requirement for carbon dioxide for their growth. Illustrative examples include the following: a. Anaerobic bacteria such as acetogens, methanogens, and propionic bacteria. b. Aerobic bacteria that degrade propane (MacMichael and Brown 1987), the branched hydrocarbon 2,6-dimethyloct-2-ene (Fall et al. 1979), or oxidize carbon monoxide (Meyer and Schlegel 1983). c. The anaerobic biotransformation of aromatic compounds may be dependent on CO2, and a review by Ensign et al. (1998) provides a brief summary of the role of CO2 in the metabolism of epoxides by Xanthobacter sp. strain Py2, and of acetone by both aerobic and anaerobic bacteria. d. The lag after diluting glucose-grown cultures of Escherichia coli into fresh medium may be eliminated by the addition of NaHCO3 and is consistent with the requirement of this organism for low concentrations of CO2 for growth (Neidhardt et al. 1974).

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BIODEGRADATION OF ENANTIOMERS The molecules of some compounds including agrochemicals such as chlorinated phenoxypropionates, certain N-substituted 2,6-dimethylanilines (e.g., metalaxyl) (Buser et al. 2002), nonplanar PCB congeners (i.e., IUPAC 91, 95, 132, 136, 149, 174, 176, 183) (Kaiser 1974), o,p′-DDT (Garrison et al. 2000), and α-hexachlorocyclohexane (Wiberg et al. 2001) are asymmetric. They may therefore exist as pairs of mirror-image enantiomers that display differential biological activity and biodegradability. Only one of them may be encountered in environmental samples, and this may plausibly be attributed to the preferential destruction or transformation of one enantiomer that is consistent with observed significant differences in the biodegradability of enantiomers. Different strategies for the biodegradation of racemates may be used and are illustrated in the following examples, and the pathways are discussed in the appropriate chapters: 1. Both enantiomers of mandelate were degraded through the activity of a mandelate racemase (Hegeman 1966), and the racemase (mdlA) is encoded in an operon that includes the following two enzymes in the pathway of degradation: S-mandelate dehydrogenase (mdlB) and benzoylformate decarboxylase (mdlC) (Tsou et al. 1990). 2. Only the R(+) enantiomer of the herbicide 2-(2-methyl-4-chlorophenoxy)propionic acid was degraded (Tett et al. 1994), although cell extracts of Sphingomonas herbicidovorans grown with the R(–) or S(–) enantiomer, respectively, transformed selectively the R(–) or S(–) substrates to 2-methyl-4-chlorophenol (Nickel et al. 1997). 3. Degradation of linear alkylbenzenes may produce chiral 3-phenylbutyrate. The degradation of the R(–) and R(–) enantiomers has been examined in Rhodococcus rhodochrous strain PB1 (Simoni et al. 1996). Whereas the S-enantiomer undergoes dioxygenation to the catechol that is a terminal metabolite, the R-enantiomer is degraded to 3-phenylpropionate followed by dioxygenation and ring fission. 4. The α-isomer of hexachlorocyclohexane exists in two enantiomeric forms, and both are degraded by Sphingomonas paucimobilis strain B90A by dehydrochlorination to 1,3,4, 6-tetrachlorocyclohexa-1,4-diene that is spontaneously degraded to 1,2,4-trichlorophenol. In this strain, there are two enzymes LinA1 and LinA2 that specifically accept (+)-αHCH or (–)-α-HCH to produce β-pentachlorocyclohexene-1 and -2. Both enzymes were 88% identical at the amino acid level, and LinA2 was identical to LinA of S. paucimobilis strain UT26. This implies that in strains such as UT26 that contain only a single dehydrochlorinase, enantioselective degradation may take place with enrichment of the nondegradable enantiomer (Suar et al. 2005). 5. Cells of Acinetobacter sp. NCIB 9871 grown with cyclohexanol carried out enantiomerically specific degradation of a racemic substituted norbornanone to a single ketone having >95% enantiomeric excess (Levitt et al. 1990). 6. Bis-(1-chloro-2-propyl)ether has two chiral centers, and exists in (R,R)-, (S,S)-, and a meso form. It is degraded by Rhodococcus sp. with a preference for the S,S-enantiomer with the intermediate formation of 1-chloro-propan-2-ol and chloroacetone (Garbe et al. 2006). This enantiomeric specificity has been of interest in other contexts, and stereospecific biotransformation has been observed. Examples include the enantiomeric oxidation of sulfides to sulfoxides (Chapter 11, Part 2) and steroid and triterpene hydroxylation (Chapter 7, Part 2). Natural systems may be quite complex. For example, the enantiomerization of phenoxyalkanoic acids containing a chiral side chain has been studied in soil using 2H2O (Buser and Müller 1997). It was shown that there was an equilibrium between the R- and S-enantiomers of both 2-(4-chloro-2methylphenoxy)propionic acid (MCPP) and 2-(2,4-dichlorophenopxy)propionic acid (DCPP) with an equilibrium constant favoring the herbicidally active R-enantiomers. The exchange reactions

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proceeded with both retention and inversion of configuration at the chiral sites. This important issue will certainly attract increasing attention in the context of the preferential microbial synthesis of intermediates of specific configuration.

SEQUENTIAL MICROBIAL AND CHEMICAL REACTIONS Microbial activity may produce reactive intermediates that undergo spontaneous chemical transformation to terminal metabolites. This is quite a frequent occurrence, and its diversity is illustrated by the following examples: 1. A bacterial strain BN6 oxidizes 5-aminonaphthalene-2-sulfonate by established pathways to 6-amino-2-hydroxybenzalpyruvate that undergoes spontaneous cyclization to 5-hydroxyquinoline-2-carboxylate (Figure 2.2a) (Nörtemann et al. 1993). 2. The oxidation of benzo[b]thiophene by strains of pseudomonads produces the sulfoxide that undergoes an intramolecular Diels–Alder reaction followed by further transformation to benzo[b]naphtho[1,2-d]thiophene (Figure 2.2b) (Kropp et al. 1994). 3. The degradation of 4-chlorobiphenyl by Sphingomonas paucimobilis strain BPSI-3 formed the intermediates 4-chlorobenzoate and 4-chlorocatechol. Fission products from the catechol reacted with NH4+ to produce chloropyridine carboxylates (Davison et al. 1996) (Figure 2.2c). 4. 4-Nitrotoluene is degraded by a strain of Mycobacterium sp. via the corresponding 4-amino3-hydroxytoluene (Spiess et al. 1998): this is dimerized abiotically to form a dihydrophenoxazinone, and after extradiol cleavage to 5-methylpyridine -2-carboxylate (Figure 2.2d). 5. The incubation of 3,5-dichloro-4-methoxybenzyl alcohol with methanogenic sludge produced the de-O-methylated compound that was transformed to 2,6-dichlorophenol, and abiotically dimerized to bis(3,5-dichloro-4-hydroxyphenyl)methane (Verhagen et al. 1998) (Figure 2.2e). 6. The transformation of aromatic amino acids to the 2-ketoacids was mediated by Morganella morganii, and these subsequently underwent a hemin-dependent chemical transformation with the production of CO (Hino and Tauchi 1987). 7. Tetrachloroethene may be degraded by bacteria via the epoxide, and chemical hydrolysis of this produces CO and CO2 from oxalyl chloride as major products, whereas only low amounts of trichloroacetate were produced (Yoshioka et al. 2002). Nitrite (or compounds at the same or lower oxidation level) is produced microbiologically from nitrate, and may then react with the substrate to produce stable end products. The production of nitrite is the sole metabolic function of the bacteria and, in view of concern over the presence of nitrate in groundwater, the following possible environmental significance of these or analogous reactions should not be overlooked: 1. A strain of Escherichia coli produces a naphthotriazole from 2,3-diaminonaphthalene and nitrite that is formed from nitrate by the action of nitrate reductase. The initial product is NO, which is converted by reactions with oxygen into the active nitrosylating agent that reacts chemically with the amine (Ji and Hollocher 1988). A comparable reaction may plausibly account for the formation of dimethylnitrosamine by Pseudomonas stutzeri during growth with dimethylamine in the presence of nitrite (Mills and Alexander 1976) (Figure 2.2f ). 2. The formation of 3,3′,4,4′-tetrachloroazobenzene, 1,3-bis(3,4-dichlorophenyl)triazine and 3,3′,4,4′-tetrachlorobiphenyl from 3,4-dichloroaniline and nitrate by E. coli plausibly involved intermediate chemical formation of the diazonium compound by reaction of the amine with nitrite (Corke et al. 1979) (Figure 2.2g).

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OH

OH

SO 2 H

CO−CO2 H

(a)

N CO2 H

NH 2

NH 2

SO

(b) S

S

S

S O

O

CO2 H

OH

N

CO2 H CO2 H

CO2 H

OH OH

Cl

Cl

Cl

(c)

CO2 H OH CHO

Cl

N

CO2 H

CH3

Cl CH3

CH3

CH3

CH 3 N CHO

(d) NO 2

NHOH

NH2

CH 3

NH 2

O

CH 3 O

N

CH2 OH

CH2 OH

CO 2 H

CO2 H

OH

CH2

Cl

CH 2

NH 2 Cl

(e) Cl

Cl

Cl

Cl

O

OH

OMe

OH

HO

Cl

Cl

NH2

N

NH2

N H

(f)

Cl

Cl

N NH2

Cl Cl

(g)

N

N

Cl

Cl

+ Cl

Cl

Cl Cl

Cl

+

Cl N

N

NH

Cl Cl

Cl Cl

FIGURE 2.2 Transformation of (a) 5-aminonaphthalene-2-sulfonate, (b) benzo[b]thiophene, (c) 4-chlorobiphenyl, (d) 4-nitrotoluene, (e) 3,5-dichlor-4-methoxybenzyl alcohol, (f) 2,3-diaminonaphthalene in presence of nitrate, and (g) 3,4-dichloroaniline presence of nitrate.

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3. Nitro-containing metabolites have been isolated from a number of substrates when the medium contained nitrate. Reduction to nitrite is the putative source of the nitro group, and two mechanisms have been suggested: (i) nitrosation by NO+ produced from HO–NO in a slightly acidic medium followed by oxidation or (ii) nitration of an intermediate arene oxide by nitrite. Since all the primary metabolites have phenolic groups and the introduced nitro groups are ortho or para to these, the first is probably the more attractive general reaction. In support of the alternative, however, an NIH shift has been demonstrated in the transformation of 2-chlorobiphenyl to 2-hydroxy-3-chlorobiphenyl by a methylotrophic organism, so that formation of an arene oxide cannot be excluded (Adriaens 1994). The following illustrate these reactions: a. The formation of nitro-containing metabolites during the degradation of 4-chlorobiphenyl by strain B-206 (Sylvestre et al. 1982). b. Corynebacterium sp. that utilizes dibenzothiophene as a sulfur source produced 2-hydroxybiphenyl, and subsequently nitrated this to form two hydroxynitrobiphenyls (Omori et al. 1992). c. The transformation of α-tocopherol by Streptomyces catenulae produced 5-nitrotocopherol in addition to quinones (Rousseau et al. 1997). d. The transformation of 2-hydroxybenz[1,4]oxazin-3-one with Gliocladium cibotii produced the intermediate 2-hydroxyacetanilide that was used to produce nitro derivatives (Zikmundová et al. 2002).

THE SPECTRUM OF ORGANISMS Bacteria in Their Natural Habitats Illustrations of the plethora of pathways used by bacteria for the degradation and biotransformation of xenobiotics are provided in Chapters 7 through 11. It is appropriate to say something here of the metabolic spectrum of specific groups of organisms, in particular, those that have hitherto achieved less prominence in discussions of biodegradation and biotransformation. The application of the methodology of molecular biology has revealed new dimensions in the range of naturally occurring organisms. Further details of some of these in the context of bioremediation are given in Chapter 13. It should be appreciated that in natural situations, bacteria may be subjected to severe nutrient limitation so that they are compelled to reproduce at extremely low rates in order to conserve their metabolic energy (Kjellberg et al. 1987; Siegele and Kolter 1992). This does not necessarily mean, however, that these organisms have negligible metabolic potential toward xenobiotics. Other slowgrowing organisms may be well adapted to the natural environment (Poindexter 1981) although they may not be numerically dominant among organisms isolated by normal procedures (Schut et al. 1993). An unusual situation has been observed in Pseudomonas putida strain mt-2 that contains the TOL plasmid pWWO. After growth with 3-methylbenzoate, cells were exposed to concentrations of toluene from 4 mg/L (growth supporting) to 130 mg/L (inhibitory) to 267 mg/L (lethal). Protein synthesis was rapidly inhibited with the concomitant production of new proteins, which were characteristic of cell starvation, and which could be suppressed by the addition of 3-methylbenzoate as carbon source. Cells exposed to 4 mg/L ceased to produce “starvation proteins” within 3 h and growth was initiated. At the higher concentrations, these proteins persisted for increasing lengths of time and at 267 mg/L there was a rapid loss of viability (Vercellone-Smith and Herson 1997). An important example is provided by Lactococcus lactis that enters a resting phase after consumption of the available carbohydrates. The ensuing noncultivable state is characterized by the synthesis of 2-methylbutyrate that is derived from l-leucine, and provides the intermediate 3-hydroxy-3-methylglutarate (Ganesan et al. 2006).

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Attention is directed to organisms that have hitherto evaded isolation or are represented by only a few cultivated examples. Such organisms may well outnumber those that have been isolated as pure cultures, frequently using elective enrichment. A few illustrations are given below: 1. Archaea belonging to the kingdom Crenarchaeota deserve attention. Although it has been assumed that these are extreme thermophiles, members of this group have been identified by molecular techniques in other habitats, for example, soils (Buckley et al. 1998), boreal forest soil (Jurgens et al. 1997), and in plant extracts (Simon et al. 2005). 2. Organisms belonging to the phylum Verrumicrobia have been detected in soil throughout the world although hitherto only a single organism Chthoniobacter flavus has been obtained in pure culture (Sangwan et al. 2004). This was made possible by the use of a dilute complex medium solidified with gellan gum in place of agar (Janssen et al. 2002), and has been extended (Sangwan et al. 2005). Before discussing some of the larger groups of microorganisms that have been implicated in biodegradation and biotransformation, brief comments are made on other groups of organisms that have hitherto attracted somewhat limited attention: a. Typically clinical organisms in environmental samples with degradative capability that include Klebsiella pneumoniae, Mycobacterium tuberculosis, and Ochrobactrum spp. b. Marine bacteria, many of which are oligotrophic, including species of Cycloclasticus, Neptunomonas, Marinobacter, and Sulfitobacter c. Gram-positive organisms belonging to the genera Mycobacterium, Rhodococcus, and Gordonia Marine and Oligotrophic Bacteria Although most illustrations have been taken from investigations of freshwater environments—lakes and rivers—fewer relate to the marine environment. In view of the area of the globe that is covered by open sea this may seem remarkable, and the degradation of xenobiotics by marine bacteria has received somewhat limited attention. Roseobacters represent one of the major groups of marine bacteria and have been found not only in a range of geographical habitats but in various degrees of association. A review provides a valuable summary, and draws attention to some of their metabolic potential including the oxidation of CO and the degradation of aromatic compounds (Buchan et al. 2005). Illustrative examples of degradation by marine bacteria are given below: 1. Strains of Marinobacter sp. that are able to degrade aliphatic hydrocarbons and related compounds have been isolated (Gauthier et al. 1992), and Marinobacter sp. strain CAB is capable of degrading 6,10,14-trimethylpentadecan-2-one under both aerobic and denitrifying conditions with the production of a range of metabolites (Rontani et al. 1997). It has been shown that Alcanivorax borkumensis is also able to degrade a number of hydrocarbons (Yakimov et al. 1998). 2. A halophilic Archaeon assigned to the genus Haloferax is able to use a restricted range of aromatic substrates including benzoate as sole sources of carbon and energy (Emerson et al. 1994). This represents an extension of the range of substrates for this genus in marine localities and has attracted attention on account of their ability to degrade polycyclic aromatic hydrocarbons (PAHs). 3. Organisms assigned to the genus Cycloclasticus have been isolated from a number of geographical locations and display a considerable metabolic versatility as follows: a. Strains were isolated from the Gulf of Mexico and from Puget Sound. In an artificial seawater medium, they were able to degrade a range of PAHs including alkylated

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4.

5.

6.

7.

8.

59

naphthalenes, phenanthrene, anthracene, and fluorene at concentrations ranging from 1 to 5 ppm (Geiselbrecht et al. 1998). Strains from both localities were numerically important, and were similar based on both 16S ribosomal DNA (rDNA) sequences and phylogenetic relationships determined from the sequences of arene dioxygenases. b. Cycloclasticus oligotrophus strain RB1 that harbors genes with a high degree of homology to that encoding xylene degradation in terrestrial pseudomonads, is able to grow at the expense of toluene, xylenes, biphenyl, naphthalene, and phenanthrene (Wang et al. 1996). c. Cycloclasticus sp. strain A5 was able to grow at the expense of naphthalenes, phenanthrenes, and fluorenes—though not anthracene (Kasai et al. 2003). Bacteria isolated from marine macrofaunal burrow sediments and assigned to Lutibacterium anuloederans were able to degrade phenanthrene in a heavily contaminated sediment (Chung and King 2001). Marine bacteria were isolated from a creosote-contaminated sediment in Puget Sound by enrichment with naphthalene (Hedlund et al. 1999). It was shown that the gene encoding the naphthalene dioxygenase ISP from this strain was not closely related to those from naphthalene-degrading strains of Pseudomonas or Burkholderia. Although analysis of 16S rDNA suggested a close relation to the genus Oceanospirillum, the differences were considered sufficient to assign these strains to a new taxon—Neptunomonas naphthovorans. Strains of Sagittula stellata and Sulfitobacter sp. that are members of the Roseobacter group of organisms were screened for the presence of 3,4-dihydroxybenzoate dioxygenase that is involved in the degradation of several aromatic carboxylates (Buchan et al. 2000). The enzyme was widely distributed in this group, and the organisms were able to degrade a number of substrates including benzoate, anthranilate, salicylate, 4-hydroxybenzoate, and vanillate. Marine roseobacters that contain bacteriochlorophyll a have been described (Oz et al. 2005), and the bacteriochlorophyll a–containing marine bacterium Porphyrobacter sanguineus was able to degrade biphenyl and dibenzofuran, though unable to use them as sole substrates for growth (Hiraishi et al. 2002). The roseobacters Silicibacter pomeroyi and Roseovarius nubinhibens were able to carry out the degradation of dimethylsulfoniopropionate to dimethylsulfide, and to methanethiol (González et al. 2003), and are discussed further in Chapter 11, Part 2.

It is necessary to take into account critical aspects of the physiology and biochemistry of these marine organisms and, although parenthetical comments on marine bacteria are made in various sections of this book, it is convenient to bring together some of their salient features. It should also be appreciated that terrestrial organisms that have a high tolerance to salinity may be isolated from inshore seawater samples. Among these are the yeasts that have been isolated from coastal marine sediments. Such organisms are excluded from the present discussion, which is restricted to oceanic water: 1. It is experimentally difficult to obtain numerical estimates of the total number of bacteria present in seawater, and the contribution of ultramicroorganisms that have a small cell volume and low concentrations of DNA may be seriously underestimated. Although it is possible to evaluate their contribution to the uptake and mineralization of readily degraded compounds such as amino acids and carbohydrates, it is more difficult to estimate their potential for degrading xenobiotics at realistic concentrations. 2. The use of conventional plating procedures may result in the isolation of only fast-growing organisms that outgrow others—which may be numerically dominant and which are unable to produce colonies on such media. The substrate concentrations used for isolation may have been unrealistically high, so that obligately oligotrophic organisms were outnumbered during attempted isolation.

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3. There is an indeterminacy in the term “oligotroph,” and the dilemma is exacerbated by the fact that it may be impossible to isolate obligate oligotrophs by established procedures. The application of DNA probes should, however, contribute to an understanding of the role of these “noncultivable” organisms. Oligotrophic bacteria in the marine environment are able to utilize low substrate concentrations, and they may be important in pristine environments. 4. During prolonged storage in the laboratory under conditions of nutrient starvation, facultatively oliogotrophic bacteria may be isolated and these display transport systems for the uptake of amino acids and glucose that are coregulated. 5. Organisms in natural ecosystems may not be actively dividing but may, nonetheless, be metabolically active. This may be particularly important for ultramicro marine bacteria in their natural habitat.

Lithotrophic Bacteria These are major groups of microorganisms that have achieved restricted prominence in discussions on biodegradation and biotransformation, and include both photolithotrophs and chemolithotrophs. Some brief comments on both are therefore justified. These organisms use CO2 as their principal, or exclusive, source of carbon, and this is incorporated into cellular material generally by the Benson–Calvin cycle. This is distinguished from heterotrophic organisms that may incorporate CO2 during degradation, for example, CO2 in the metabolism of epoxides by Xanthobacter sp. strain Py2, and of acetone by both aerobic and anaerobic bacteria. Energy for growth is obtained either from photochemical reactions (photolithotrophs), or by chemical oxidation of inorganic substrates such as reduced forms of nitrogen or sulfur (chemolithotrophs). In some organisms, organic carbon can be taken up and incorporated during growth even in organisms that are obligately chemolithotrophic or photolithotrophic. Organic carbon may, however, have an inhibitory effect on growth. Some species and strains of these organisms may also grow heterotrophically using organic carbon as sources of both energy and cell carbon. Attention is directed to reviews that cover the sometimes controversial aspects of lithotrophy and autotrophy (Kelly 1971; Rittenberg 1972; Whittenbury and Kelly 1977; Matin 1978). It is worth noting that many aerobic bacteria that belong to groups with well-established heterotrophic activity are also chemolithoautotrophic and use the oxidation of hydrogen as their source of energy (Bowien and Schlegel 1981). Attention is drawn to the reassignment of the Pseudomonas strains P. flava, P. pseudoflava, and P. palleroni to the genus Hydrogenophaga, and the degradative activity of species of Xanthobacter. Attention has been directed to obligate lithotrophs such as those that oxidize carbon monoxide or hydrogen. Analysis of DNA extracts and PCR amplification for the cbbL gene that encodes the large subunit of ribulose-1,5-bisophosphate carboxylase/oxygenase that is typically found in obligate lithotrophs was carried out. The PCR products were cloned, sequenced, and analyzed, and the results showed that populations depended on the plant cover and land use, although the relevant populations of ammonia-oxidizing organisms were small in comparison with those of facultative lithotrophs (Tolli and King 2005). Anaerobic lithotrophs that oxidize ammonium using nitrite as electron acceptor (anammox) are noted later in the section on anaerobic bacteria. It has become increasingly clear that representatives of chemolithotrophic microorganisms may be effective in carrying out the transformation of xenobiotics. As illustration, attention is directed to three groups of organisms: (1) the ammonia-oxidizers; (2) the thiobacilli; and (3) algae and cyanobacteria. Habitats to which these organisms are physiologically adapted should therefore be considered in discussions on biodegradation and biotransformation.

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Phototrophic Organisms The metabolic significance of oxygenic algae and cyanobacteria has received relatively limited attention in spite of the fact that they are important components of many ecosystems and may, for example, in the marine environment, be of primary significance. Whereas the heterotrophic growth of algae at the expense of simple carbohydrates, amino acids, lower aliphatic carboxylic acids, and simple polyols is well documented (Neilson and Lewin 1974), the potential of algae for the metabolism of xenobiotics has been much less extensively explored. Among these metabolic possibilities, which have received less attention than they deserve, the following are used as illustration: 1. The transformation—though not apparently the degradation—of naphthalene has been examined in cyanobacteria and microalgae, including representatives of green, red, and brown algae, and diatoms (Cerniglia et al. 1980a, 1982). The transformation of biphenyl (Cerniglia et al. 1980b), aniline (Cerniglia et al. 1981), and methylnaphthalenes (Cerniglia et al. 1983) has been examined in cyanobacteria (Figure 2.3). Phenanthrene is metabolized by Agmenellum quadruplicatum to the trans-9,10-dihydrodiol by a monooxygenase system, and the 1-hydroxyphenanthrene that was formed transiently was O-methylated (Narro et al. 1992). The biotransformation of benzo[a]pyrene has been demonstrated in a number of green algae, though this was not metabolized by a chlamydomonad, a chrysophyte, a euglenid, or a cyanobacterium (Warshawsky et al. 1995). The relative amounts of the products depended on the light sources and their intensity, and included 9,10-, 4,5-, 11,12-, and 7,8-dihydrodiols, the toxic 3,6-quinone, and phenols. The 11,12-, 7,8-, and 4,5-dihydrodiols produced by Selenastrum capricornutum had the cis configuration, which suggests that their formation was mediated by a dioxygenase (Warshawsky et al. 1988). It is worth noting that Ochromonas danica is able to degrade phenol by extradiol fission of the initially formed catechol (Semple and Cain 1996). 2. Some green algae are able to use aromatic sulfonic acids (Figure 2.4a) (Soeder et al. 1987) and aliphatic sulfonic acids (Figure 2.4b) (Biedlingmeier and Schmidt 1983) as sources of sulfur. Cultures of Scenedesmus obliquus under conditions of sulfate limitation metabolized naphthalene-1-sulfonate to 1-hydroxy-naphthalene-2-sulfonate and the glucoside of naphth-1-ol (Kneifel et al. 1997). These results are consistent with formation of a 1,2-epoxide followed by an NIH shift. 3. The cyanobacteria Anabaena sp. strain PCC 7120 and Nostoc ellipsosporum dechlorinated γ-hexachloro[aaaeee]cyclohexane in the light in presence of nitrate to γ-pentachlorocyclohexene (Figure 2.5), and to a mixture of chlorobenzenes (Kuritz and Wolk 1995). The reaction is dependent on the functioning of the nir operon involved in nitrite reduction (Kuritz et al. 1997). OH

(a)

(b)

OH

FIGURE 2.3 Examples of the biotransformation of aromatic compounds by algae.

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Environmental Degradation and Transformation of Organic Chemicals SO 3 H (b) CH 3 −CH 2 −SO 3 H

(a)

FIGURE 2.4

CH 3 −(CH 2 )1 2 −CH

CO 2 CH 3 SO 3 H

Organosulfur compounds used as sources of sulfur by algae. Cl

Cl

Cl Cl

Cl

Cl

Cl

Cl Cl

Cl

Cl

OH

Cl Cl Cl

HO

Cl

OH

Cl OH

Cl

Cl

Cl

Cl

Cl

Cl

FIGURE 2.5 Transformation of γ-hexachlorocyclohexane by cyanobacteria.

4. Representatives of the major groups of algae are able to use a range of amino acids as nitrogen sources (Neilson and Larsson 1980). 5. The transformation of DDT to DDE—albeit in rather low yield—by elimination of one molecule of HCl has been observed in several marine algae (Rice and Sikka 1973). There has been a revival of interest in the metabolic potential of anaerobic phototrophic bacteria, particularly, the purple nonsulfur organisms that can degrade aromatic compounds (Khanna et al. 1992). Such organisms are widely distributed in appropriate ecosystems, and may therefore play a significant role in the degradation of xenobiotics. Less appears to be known of the potential of other anaerobic phototrophs such as the purple and green sulfur bacteria to degrade xenobiotics. Dehalogenation of a number of halogenated alkanoic acids has been observed with Rhodospirillum rubrum, R. photometricum, and Rhodopseudomonas palustris. Substrates include 2- and 3-chloropropionic acid for all of these organisms, chloroacetate for R. photometricum, and 2-bromopropionate for R. rubrum and Rh. palustris (McGrath and Harfoot 1997). Fe(II) (Ehrenreich and Widdel 1994) and taurine (Novak et al. 2004) may serve as electron donors for the anaerobic phototrophic growth of purple nonsulfur bacteria. Aerobic and Facultatively Anaerobic Bacteria The well-established metabolic versatility of groups such as the pseudomonads and their numerous relatives, and the methanotrophs has possibly deflected attention from other groups that may be present in aquatic systems and which may play an important role in determining the fate of xenobiotics. This is increasingly being rectified with the isolation and description of new taxa. Although the potential of other Gram-negative groups including the pseudomonads, acinetobacters, moraxellas, and species of Alcaligenes is well established, Gram-positive groups seem to have

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achieved generally somewhat less prominence in aquatic systems. In the succeeding paragraphs, some examples of the metabolic importance of a few of these groups of organisms are presented. Gram-Positive Aerobic Bacteria The metabolic versatility of organisms belonging to the genera Mycobacterium and Rhodococcus is becoming well established. A taxonomic word of caution is appropriate in view of the historical importance of similar organisms that were designated as mycobacteria (Gray and Thornton 1928), although they were not acid-fast, and some at least would be currently assigned to the genus Rhodococcus. This illustrates a potentially serious taxonomic pitfall that should be avoided, since the distinction between the genera Mycobacterium and Rhodococcus has not always been unequivocal in the older literature (Finnerty 1992). Indeed, the organism that is discussed in Chapter 9, Part 2, and initially described as Rhodococcus chlorophenolicus has now been transferred to the genus Mycobacterium (Häggblom et al. 1994; Briglia et al. 1994). Conversely, an organism capable of degrading acetylene and assigned (DeBont et al. 1980) to the genus Mycobacterium almost certainly belongs to the genus Rhodococcus. There has been increasing interest in both mycobacteria and rhodococci. This has been sustained by the potential application of rhodococci in biotechnology (Finnerty 1992) since they have a broad metabolic potential. The analysis of fatty acids, mycolic acids, and diaminopimelic acid has been a valuable chemotaxonomic development, in addition to the use of sequencing. Mycobacteria Alkanes and Haloalkanes

Mycobacteria can oxidize short-chain alkenes (DeBont et al. 1980), and the organism known under the invalid specific name “Mycobacterium paraffinicum” (Wayne et al. 1991) was able to degrade a number of alkanes. It has also been shown that strains of mycobacteria grown with propane are able to oxidize the apparently unrelated substrate trichloroethene (Wackett et al. 1989). Strains of mycobacteria have been shown to degrade a range of halogenated alkenes. These include (a) vinyl chloride (Hartmans and de Bont 1992; Coleman and Spain 2003) and (b) Mycobacterium sp. strain GP1 that degrades 1,2-dibromoethane by a pathway that avoids the production of the toxic 2-bromoethanol and 2-bromoacetaldehyde by formation of the epoxide (Poelarends et al. 1999). Possibly, more remarkable is the metabolic capacity of species of mycobacteria including the human pathogen M. tuberculosis strain H37Rv (Jesenská et al. 2000). On the basis of amino acid and DNA sequences, the strain that was used contained three halohydrolases and the activity of the dehalogenase for debromination of 1,2-dibromoethane by other species of mycobacteria is given in Table 2.1. The haloalkane dehalogenase gene from M. avium has been cloned and partly characterized (Jesenská et al. 2002). PAHs

There has been a revival of interest in the role of mycobacteria in the degradation of polycyclic hydrocarbons including naphthalene (Kelley et al. 1990), phenanthrene (Guerin and Jones 1988), anthracene (van Herwijnen et al. 2003), and pyrene (Heitkamp et al. 1988; Grosser et al. 1991; TABLE 2.1 Specific Activity (µmol Bromide Produced per mg Protein per min) of Dehalogenase from Selected Species of Mycobacteria toward 1,2-dibromoethane Taxon M. bovis BCG MU10 M. fortuitum MU8 M. triviale MU3 M. smegmatis CCM4622

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Activity

Taxon

Activity

99 76 61 49

M.avium MU1 M. phlei CCM 5639 M. parafortuitun MU2 M. chelonae

36 22 22 20

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TABLE 2.2 Degradation of PAHs by Mycobacteria Substrate Nap, Phe, Flu, 3-Me-Chol Phe, Pyr, Flu Pyr, BaPyr, BaAnth Nap, Anth, Phe, Flu, Pyr, BaAnth, BaPyr

Organism Mycobacterium sp. Mycobacterium sp. BB1 Mycobacterium sp. RJGII-135 Mycobacterium vanbaalenii PYR-1

Reference Heitkamp et al. (1988) Boldrin et al. (1993) Schneider et al. (1996) Kim et al. (2006)

Note: Bi, biphenyl; Nap, naphthalene; Phe, phenanthrene; Anth, anthracene; Flu, fluoranthene; Pyr, pyrene; Chr, chrysene; BaAnth, benz[a]anthracene; BaPyr, benzo[a]pyrene; BbFlu, benzo[b]fluoranthene; DBaAnth, dibenz[a,h]anthracene; 3-Me-Chol, 3-methylcholanthrene.

Krivobok et al. 2003) (Table 2.2). Both of the pyrene-degrading strains belong to the group of fastgrowing scotochromogenic mycobacteria (Govindaswami et al. 1995). A strain of Mycobacterium sp. that is able to use all these substrates as a sole source of carbon and energy has also been isolated (Boldrin et al. 1993). Halogenated Phenols

Considerable interest has been expressed in the chlorophenol-degrading organism Mycobacterium chlorophenolicum (R. chlorophenolicus) (Apajalahti et al. 1986), partly motivated by its potential for application to bioremediation of chlorophenol-contaminated industrial sites (Häggblom and Valo 1995). Nitroarenes

The pathway used by a strain of Mycobacterium sp. for the degradation of 4-nitrotoluene is initiated by reduction to the hydroxylamine and rearrangement to 3-hydroxy-4-aminotoluene before further degradation (Spiess et al. 1998). Morpholine

This can be degraded by several mycobacteria including Mycobacterium aurum strain MO1 (Combourieu et al. 1998), Mycobacterium strain RP1 (Poupin et al. 1998), and Mycobacterium chelonae (Swain et al. 1991). The reaction is initiated by a cytochrome P450 monooxygenase that is also active against pyrrolidine and piperidine (Poupin et al. 1998). Carbon Monoxide

The degradation of carbon monoxide has attracted recent attention, and a strain of mycobacterium is able to degrade this unusual substrate (Park et al. 2003). Rhodococci

Only a selection of examples is given below to illustrate the metabolic versatility of rhodococci, and further details are given in the appropriate sections of Chapters 7 through 11: 1. A strain of Rhodococcus sp. was capable of degrading a number of chlorinated aliphatic hydrocarbons including vinyl chloride and trichloroethene, as well as the aromatic hydrocarbons benzene, naphthalene, and biphenyl (Malachowsky et al. 1994). 2. A strain of Rhodococcus opacus isolated by enrichment with chlorobenzene was able to grow at the expense of a wide range of halogenated compounds. These included 1,3- and 1,4-dichlorobenzene, 1,3- and 1,4-dibromobenzene, 2-, 3-, and 4-fluorophenol, 2-, 3-, and 4-chlorophenol, 4-nitrophenol, 3- and 4-fluorobenzoate, and 3-chlorobenzoate (Zaitsev et al. 1995). 3. Several rhodococci have attracted interest for their ability to degrade PCBs including an organism (Acinetobacter sp. strain P6) now assigned to Rhodococcus globerulus (Asturias

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and Timmis 1993), R. erythropolis (Maeda et al. 1995), and Rhodococcus sp. strain RHA1 (Seto et al. 1995). 4. The degradation of 2,4,6-trichlorophenol by Rhodococcus percolatus (Briglia et al. 1996). 5. 2,4,6-Trinitrophenol is degraded in a reaction involving ring reduction by hydride transfer from an NADPH-dependent F420 reductase (Hofmann et al. 2004). 6. The degradation of cyclopropanecarboxylate by Rhodococcus rhodochrous (Toraya et al. 2004). 7. The degradation of isoprene by Rhodococcus sp. strain AD45 involves a glutathionemediated reaction (van Hylckama Vlieg et al. 2000) and is noted again in Chapter 7, Part 1. 8. A number of aliphatic ethers can be degraded by Rhodococcus sp. strain DEE 5151 (Kim and Engesser 2004). A number of biotransformations have been accomplished by rhodococci, including, for example, the hydrolysis of nitriles including polyacronitriles (Tauber et al. 2000), and the reduction of the conjugated C=C double bond in 2-nitro-1-phenylprop-1-ene (Sakai et al. 1985). Other Gram-Positive Organisms Gordonia

A review has summarized the taxonomic status of the genus Gordonia (Arenskötter et al. 2004), and drawn attention to important degradations carried out by species of the genus. These include benzothiophene by G. desulfuricans, dibenzothiophene by G. amicalis, 3-alkylpyridines by G. nitida, hexadecane by G. alkanivorans, 2,4-dinitrophenol by G. koreensis, and cis-1,4-polyisoprene by G. polyisoprenivorans (references in Braaz et al. 2004) that is noted in Chapter 3, Part 1 on dioxygenases. Actinomycetes

The role of cytochrome P450 oxygenations in actinomycetes has been reviewed (O’Keefe and Harder 1991) and, although these organisms have been exhaustively explored as the source of clinically valuable antibiotics, their degradative capability has been less extensively examined. It has been shown that they have a virtually ubiquitous occurrence in tropical and subtropical marine sediments (Mincer et al. 2002), and all of these have been assigned to two taxa Salinispora arenicola and S. tropica (Mincer et al. 2005). Their occurrence has been explored in the marine environment as possible sources of biologically active metabolites (Magarvey et al. 2004). Streptomyces

These have been examined for the degradation of a number of contaminants, and two pathways for the introduction of oxygen have been demonstrated: 1. Cytochrome P450 monooxygenation a. The transformation by Streptomyces griseus of benzene, toluene, naphthalene, biphenyl, and benzo[a]pyrene to the corresponding phenols has been observed (Trower et al. 1989). The oxidation of phenanthrene to (–)trans-[9S,10S]-9,10-dihydrodihydroxyphenanthrene by Streptomyces flavovirens with minor amounts of 9-hydroxyphenanthrene (Sutherland et al. 1990) is plausibly carried out by epoxidation and epoxide hydrolysis rather than by typical dioxygenation. b. The transformation of sulfonylureas by Streptomyces griseolus grown in a complex medium containing glucose when the methyl group of the heterocyclic moieties is hydroxylated and, for some substrates subsequently oxidized to the carboxylic acid (Romesser and O’Keefe 1986; O’Keefe et al. 1988).

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2. Dioxygenation a. The metabolism of aromatic substrates has been examined in Streptomyces setonii, and the degradation of vanillate involved in decarboxylation to guaiacol followed by intradiol fission of the catechol that was produced (Pometto et al. 1981). This strain could also metabolize cinnamate and related carboxylic acids (Sutherland et al. 1983), and strains of Amycolatopsis sp. and Streptomyces sp. were able to metabolize benzoate and salicylate that were degraded by dioxygenation (Grund et al. 1990). b. The degradation of 4-hydroxybenzoate and vanillate has been described for Streptomyces sp. strain 2065, and the pathway involving intradiol dioxygenation of the initially formed 3,4-dihydroxybenzoate has been characterized (Iwagami et al. 2000). c. The degradation of 2,4,6-trichlorophenol has been described for Streptomyces rochei strain 303, and the dioxygenase that brings about intradiol fission of the initially formed 2,6-dichlorohydroquinone has been characterized (Zaborina et al. 1995). d. The degradation of cis-poly-1,4-isoprene has been examined in Streptomyces coelicolor 1A, and a number of oxidation products identified (Bode et al. 2000). These were produced by oxidations of the subterminal double bonds followed by β-oxidation, hydroxylation, oxidation, and fission of the 3-ketoacids. Gram-Positive Facultatively Anaerobic Organisms The genus Staphylococcus is traditionally associated with disease in humans. The demonstration (Monna et al. 1993) that a strain of Staphylococcus auriculans—isolated by enrichment with dibenzofuran and with no obvious clinical association—could degrade this substrate and carry out limited biotransformation of fluorene and dibenzo[1,4]dioxin serves to illustrate the unsuspected metabolic potential of facultatively anaerobic Gram-positive organisms. A satisfying evaluation of the metabolic potential of microorganisms in natural ecosystems should not therefore fail to consider all these organisms, which are certainly widespread, and to distinguish between rates of degradation and metabolic potential: slow-growing organisms may be extremely important in degrading xenobiotics in natural ecosystems. Gram-Negative Aerobic Bacteria A taxonomic note: there have been substantial developments in the taxonomy of pseudomonads, and many new genera have been proposed including, for example, Sphingomonas, Comamonas, and Variovorax, while denitrifying organisms described as pseudomonads have been referred to the general Thauera and Azoarcus (Anders et al. 1995). Attention is briefly drawn to groups that are not only widely distributed, but which have also previously attracted attention primarily in the context of nitrogen fixation. 1. Azotobacters. Burk and Winogradsky in the 1930s showed that these could readily be obtained from soil samples by elective enrichment with benzoate. The degradative pathway for benzoate has been elucidated (Hardisson et al. 1969), and the range of substrates extended to 2,4,6-trichlorophenol (Li et al. 1992; Latus et al. 1995). The enzyme from Azotobacter sp. strain GP1 that catalyzes the formation of 2,6-dichlorohydroquinone from 2,4,6-trichlorophenol is a monooxygenase that requires NADH, FAD, and O2 (Wieser et al. 1997) and, in the absence of a substrate, results in unproductive formation of H2O2. It is also able to accept other chlorophenols with the consumption of NADH including 2,4-, 2,6-, 3,4-dichloro-, 2,4,5-trichloro-, and 2,3,4,5- and 2,3,4,6-tetrachlorophenol. 2. Rhizobia. Taxa belonging to both the genera Rhizobium and Bradyrhizobium are capable of degrading simple aromatic compounds including benzoate (Chen et al. 1984) and 4-hydroxybenzoate (Parke and Ornston 1986; Parke et al. 1991). It has been shown that 4-hydroxybenzoate hydroxylase is required for the transport of 4-hydroxybenzoate into the cell (Wong et al. 1994). In strains of Rhizobium trifolium, the metabolism of benzoate involves either 3,4-dihydroxybenzoate (protocatechuate) 3,4-dioxygenase (Chen et al. 1984), or catechol

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TABLE 2.3 The Major Genera within the Former Genus Sphingomonas and Their Role in Degradation Genus Sphingobium

Novosphingobium Sphingopyxis

Species herbicidovorans chlorophenolicus yanoikuyae indicum japonicum francense aromaticivorans subarctica macrogolitabida

Degradation Dichlorophenoxypropionate Chlorophenols Aromatic hydrocarbons Hexachlorocyclohexanes Hexachlorocyclohexanes Hexachlorocyclohexanes Aromatic hydrocarbons Chlorophenols Polyethylene glycol

1,2-dioxygenase (Chen et al. 1985). The degradation of the toxic N-alkylpyrid-4-one mimosine by Rhizobium sp. strain TAL1145 has been described (Awaya et al. 2005). Rhizobia have quite a broad degradative capability that is illustrated by the following examples: a. PCB congeners (Damaj and Ahmad 1996) b. Flavones (Rao et al. 1991; Rao and Cooper 1994) c. Dechlorination—though not the degradation—of atrazine (Bouquard et al. 1997) d. Phosphonomycin (1,2-epoxypropylphosphonate) as a source of carbon, energy, and phosphorus (McGrath et al. 1998) 3. Many organisms with degradative capability have been assigned to species in the genus Sphingomonas. These include the degradation of chlorophenols, aromatic hydrocarbons, 6-aminonaphthalene-2-sulfonate, and nonylphenol. The description of the genus Sphingomonas has been revised, and separated into three genera, Sphingobium, Novosphingobium, and Sphingopyxis (Table 2.3) (Takeuchi et al. 2001). Important assignments to these new genera include a number of degradations, and a strain of Novosphingobium sp. represented the dominant organisms that degraded chlorophenols in a groundwater remediation system (Tiirola et al. 2002). Gram-Negative Facultatively Anaerobic Bacteria Enterobacteriaceae

Although facultatively anaerobic bacteria and especially those belonging to the family Enterobacteriaceae have a long history as the agents of disease in man, there is increasing evidence for their importance in a wide range of environmental samples. Methods for their identification and classification have been extensively developed and have traditionally used the ability to ferment a wide range of carbohydrates as taxonomic characters. This has had the possibly unfortunate effect of deflecting attention from the capability of these organisms to degrade other classes of substrates. Their ability to utilize, for example, 3- and 4-hydroxybenzoates (Véron and Le Minor 1975), and nicotinate (Grimont et al. 1977) under aerobic conditions has, however, been quite extensively used for taxonomic classification. Some examples used to illustrate the metabolic capabilities of these somewhat neglected organisms that include species of the genera Citrobacter, Enterobacter, Hafnia, Klebsiella, Salmonella, and Serratia, as well as more recently described genera, are given below: 1. The degradation of DDT by organisms designated Aerobacter aerogenes (possibly Klebsiella aerogenes) (Wedemeyer 1967) (Figure 2.6), and the partial reductive dechlorination of methoxychlor by K. pneumoniae (Baarschers et al. 1982). 2. The biotransformation of methyl phenyl phosphonate to benzene by K. pneumoniae (Cook et al. 1979) (Figure 2.7a). Further examples of the cleavage of the carbon–phosphorus

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Environmental Degradation and Transformation of Organic Chemicals CCl3

Cl

CHCl2

CH

FIGURE 2.6

Cl

Cl

CCl2

CH

Cl

Cl

C

Cl

Degradation of DDT by Aerobacter aerogenes. Cl

PO−OMe

Cl

Cl

Cl

Cl

Cl

Cl

Cl

(b)

(a)

Cl

Cl

CO 2 H CO 2 H

CN

(d)

(c) Br

Br OH

Br

Br OH

OMe OH

OMe OH

FIGURE 2.7 Transformation of (a) methylphenyl phosphonate, (b) γ-hexachlorocyclohexane, (c) 3,5-dibromo4-hydroxybenzonitrile, and (d) decarboxylation.

3. 4. 5.

6.

7.

8.

9. 10.

bond by other members of the Enterobacteriaceae, and a discussion of the metabolism of phosphonates is given in Chapter 11, Part 4. The biotransformation of 2,4,6-trihydroxy-1,3,5-triazine and atrazine under anaerobic conditions by an unidentified facultative anaerobe (Jessee et al. 1983). The biotransformation of γ-hexachloro[aaaeee]cyclohexane to tetrachlorocyclohexene by Citrobacter freundii (Figure 2.7b) (Jagnow et al. 1977). The biotransformation of 3,5-dibromo-4-hydroxybenzonitrile to the corresponding acid by a strain of K. pneumoniae ssp. ozaenae which uses the substrate as sole source of nitrogen (Figure 2.7c) (McBride et al. 1986). The decarboxylation of 4-hydroxycinnamic acid to 4-hydroxystyrene, and of ferulic acid (3-methoxy-4-hydroxycinnamic acid) to 4-vinylguaiacol by several strains of Hafnia alvei and H. protea, and by single strains of Enterobacter cloacae and K. aerogenes (Figure 2.7d) (Lindsay and Priest 1975). The decarboxylase has been purified from Bacillus pumilis (Degrassi et al. 1995). Several taxa of Enterobacteriaceae including Morganella morganii, Proteus vulgaris, and Raoultella (Klebsiella) planticola are able to decarboxylate the amino acid histidine, which is abundant in the muscle tissue of scombroid fish (Yoshinaga and Frank 1982; Takahashi et al. 2003). The histamine produced has been associated with an incident of scombroid fish poisoning (Taylor et al. 1989). The metabolism of ferulic acid (3-methoxy-4-hydroxycinnamic acid) by Ent. cloacae to a number of products including phenylpropionate and benzoate (Figure 2.8) (Grbic-Galic 1986). Utilization of uric acid as a nitrogen source by strains of Aer. aerogenes, K. pneumoniae, and Serratia kiliensis (Rouf and Lomprey 1968). The sequential reduction of one of the nitro groups of 2,6-dinitrotoluene by Salmonella typhimurium (Sayama et al. 1992)—a taxon that is not generally noted for its metabolic activity.

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69 CO 2 H CH2−CH2−CO2H

CO 2 H

+ OMe OH

FIGURE 2.8

Metabolism of ferulic acid by Enterobacter cloacae.

11. Under conditions of oxygen limitation, strains of Morganella morganii and Providencia rettgeri degraded hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) after initial reduction to the nitroso compounds (Kitts et al. 1994). 12. After growth in a medium containing suitable reductants such as glucose, a strain of Enterobacter agglomerans was able to reduce tetrachloroethene successively to trichloroethene and cis-1,2- dichloroethene (Sharma and McCarty 1996). 13. Degradation of pentaerythritol tetranitrate by Ent. cloacae (French et al. 1996) and of glyceryltrinitrate by K. oxytoca (Marshall & White 2001). 14. Degradation by Enterobacter sp. of chlorpyrifos with the formation of 3,5,6-pyridin-2-one and diethylthiophosphate that was used for growth and energy (Singh et al. 2004). 15. Degradation of naphthalene, fluoranthene, and pyrene by Leclercia adecarboxylata that has hitherto been of clinical origin (Sarma et al. 2004). 16. RDX and its partial reduction product hexahydro-1-nitroso-3,5-dinitro-1,3,5-triazine were degraded by K. pneumoniae to methylenedinitramine, and then to CH 2O and methanol, while abiotic reactions produced N2O (Zhao et al. 2002). 17. The taxonomic application of the ability of enteric organisms to grow with 4-hydroxyphenylacetate (Cooper and Skinner 1980) and 3-hydroxyphenylpropionic acid (Burlingame and Chapman 1983) has been established. In addition, it has been demonstrated that the enzyme that carries out the hydroxylation has a wide substrate range extending to 4-methylphenol, and even to 4-chlorophenol (Prieto and Garcis 1994). Vibrionaceae

Only a few examples have emerged of the degradative capability of these organisms, and include the following: a. An organism, tentatively identified as a strain of Vibrio sp., was able to degrade the 2-carboxylates of furan, pyrrole, and thiophene (Evans and Venables 1990). b. A strain of Aeromonas sp. was able to degrade phenanthrene through o-phthalate (Kiyohara et al. 1976).

BACTERIAL METABOLISM OF C1 COMPOUNDS: METHANOTROPHS, METHYLOTROPHS, AND RELATED ORGANISMS Methane Monooxygenase and Related Systems A few of the reactions carried out by the monooxygenase system of methanotrophic bacteria are summarized in Figure 2.9, and it is on account of this that methylotrophs have received attention for their technological potential (Lidstrom and Stirling 1990). An equally wide metabolic potential has also been demonstrated for cyclohexane monooxygenase, which has been shown to accomplish two broad types of reaction: one in which formally nucleophilic oxygen reacts with the substrate, and

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Environmental Degradation and Transformation of Organic Chemicals CH 3 (CH 2 )6 −CH 2OH

CH3(CH2)6−CH3

CH 3 −Br

O

CH 3 (CH 2 )5 −CH(OH)CH 3

;

CH 2 O

;

CH 3 −CH

CH 3 CH 2 −O−CH 2 CH 3

CH 2

CH 3 −CH

CH 2

CH 3CH 2 OH

OH

; N

N

CH 3

CH 3

O

CH 2OH

CO 2H

OH

FIGURE 2.9

Reactions mediated by the monooxygenase system of methanotrophic bacteria.

CH 2 −CO−CH 3

S

CH2−O−CO−CH3

O

FIGURE 2.10

S

S O

O

Reactions mediated by cyclohexane monooxygenase.

the other in which formally electrophilic oxygen is involved (Figure 2.10) (Branchaud and Walsh 1985). In addition, it has emerged that the monooxygenase system in methanotrophs is similar to that in the nitrite-oxidizing bacteria, and that the spectrum of biotransformations is equally wide. There are also important groups of facultatively C1-utilizing bacteria: (a) members of the genus Xanthobacter (Padden et al. 1997), and (b) Methylocella silvestris that was isolated from tundra soils and had the unusual capacity to use not only methane and methanol as energy sources, but also C2 and C3 substrates such as acetate and malate (Dedysh et al. 2005). The following illustrate some of the biotransformations that have been observed with Nitrosomonas europaea. These are particularly interesting since this organism has an obligate

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71 OH

OH

(a)

OH

HO 2 C

N

Cl

Cl2 CH

N

Cl

(b) Cl3 C

N

Cl

FIGURE 2.11 Metabolism by Nitrosomonas europaea (a) benzene, (b) 2-chloro-6-trichloromethylpyridine.

dependency on CO2 as carbon source, and has traditionally been considered to be extremely limited in its ability to use organic carbon for growth or transformation: 1. The oxidation of benzene to phenol and 1,4-dihydroxybenzene (Figure 2.11a) (Hyman et al. 1985), both side chain and ring oxidation of ethyl benzene, and ring-hydroxylation of halogenated benzenes and nitrobenzene (Keener and Arp 1994). 2. The oxidation of alkanes (C1–C8) to alkanols, and alkenes (C2–C5) to epoxides (Hyman et al. 1988). 3. The oxidation of methyl fluoride to formaldehyde (Hyman et al. 1994), and of chloroalkanes at carbon atoms substituted with a single chlorine atom to the corresponding aldehyde (Rasche et al. 1991). 4. The oxidation of a number of chloroalkanes and chloroalkenes including dichloromethane, chloroform, 1,1,2-trichloroethane, and 1,2,2-trichloroethene (Vannelli et al. 1990). Although the rate of cometabolism of trihalomethanes increased with levels of bromine substitution so also did toxicity. Both factors must therefore be evaluated in the possible application of this strain (Wahman et al. 2005). 5. The oxidation of the trichloromethyl group in 2-chloro-6-trichloromethylpyridine to the corresponding carboxylic acid occurs at high oxygen concentrations during cooxidation of ammonia or hydrazine. At low oxygen concentrations, however, in the presence of hydrazine, reductive dechlorination to 2-chloro-6-dichloromethylpyridine occurs (Vannelli and Hooper 1992) (Figure 2.11b). 6. A range of sulfides including methylsulfide, tetrahydrothiophene, and phenylmethylsulfide are oxidized to the corresponding sulfoxides (Juliette et al. 1993). As for methanotrophic bacteria, such transformations are probably confined neither to a single organism nor to strains of specific taxa within the group. For example, both Nitrosococcus oceanus and Nitrosomonas europaea are able to oxidize methane to CO2 (Jones and Morita 1983; Ward 1987). The versatility of this group of organisms clearly motivates a reassessment of their ecological significance particularly in the marine environment where they are widely distributed. The overall similarity of these transformations to those carried out by eukaryotic cytochrome P450 systems (Guengerich 1990) is striking.

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ANAEROBIC BACTERIA There are a number of reasons for the increased interest in transformations carried out by anaerobic bacteria: a. After discharge into the aquatic environment, many xenobiotics are partitioned from the aquatic phase into the sediment phase. b. Sediments in the vicinity of industrial discharge often contain readily degraded organic matter, and the activity of aerobic and facultatively anaerobic bacteria then renders these sediments effectively anaerobic. The fate of xenobiotics in many environments is therefore significantly determined by the degradative activity of anaerobic bacteria. The terms anaerobic and anoxic are purely operational, and imply merely the absence of air or oxygen, and the absolute distinction between aerobic and anaerobic organisms is becoming increasingly blurred. The problem of defining anaerobic bacteria may therefore best be left to philosophers. Possibly, the critical issue is the degree to which low concentrations of oxygen are either necessary for growth or toxic. During the growth of bacteria in the absence of externally added electron acceptors, the term fermentation implies that a redox balance is achieved between the substrate (which may include CO2) and its metabolites. A few following examples are given to illustrate the apparently conflicting situations that may be encountered, and the gradients of response that may be elicited: 1. Although strictly anaerobic bacteria do not generally grow in the presence of high-potential electron acceptors such as oxygen or nitrate, an intriguing exception is provided by an obligately anaerobic organism that uses nitrate as electron acceptor during the degradation of resorcinol (Gorny et al. 1992). This isolation draws attention to the unknown extent to which such organisms exist in natural systems, since strictly anaerobic conditions are not always used for the isolation of organisms using nitrate as electron acceptor. Other formally similar organisms are either facultatively anaerobic or have a fermentative metabolism. In contrast, the normally strictly anaerobic sulfate-reducing organisms Desulfobulbus propionicus and Desulfovibrio desulfuricans may grow by reducing nitrate or nitrite to ammonia using hydrogen as electron donor (Seitz and Cypionka 1986). Desulfovibrio vulgaris Hildeborough is capable of growing at oxygen concentrations of 0.24–0.48 µM, and indeed it was suggested that this organism might protect anoxic environments from adverse effects resulting from intrusion of oxygen (Johnson et al. 1997). In the presence of nitrate, the acetogen Clostridium thermoaceticum oxidizes the O-methyl groups of vanillin or vanillate to CO2 without production of acetate, which is usually formed in the absence of nitrate (Seifritz et al. 1993). 2. Some anaerobic organisms such as clostridia are appreciably tolerant of exposure to oxygen, whereas others such as Wolinella succinogenes that have hitherto been classified as anaerobes are in fact microaerophilic (Han et al. 1991). However, organisms such as Nitrosomonas europaea that normally obtain energy for growth by the oxidation of ammonia to nitrite, may apparently bring about denitrification of nitrite under conditions of oxygen stress (Poth and Focht 1985), or under anaerobic conditions in the presence of pyruvate (Abeliovich and Vonshak 1992). 3. It is important to note the existence of microaerotolerant or microaerophilic organisms. The example of W. succinogenes has already been noted, and another is provided by Malonomonas rubra that uses malonate as sole source of carbon and energy (Dehning and Schink 1989a). On the other hand, Propionigenium modestum that obtains its energy by the decarboxylation of succinate to propionate is a strictly anaerobic organism (Schink and Pfennig 1982).

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Attention is directed to two groups of anaerobic bacteria that display metabolic versatility toward structurally diverse substrates—clostridia and anaerobic sulfate reducers. Clostridia The classical studies on the anaerobic metabolism of amino acids, purines, and pyrimidines by clostridia not only set out the relevant experimental procedures and thereby laid the foundations for virtually all future investigations, but also brought to light the importance and range of coenzyme B12mediated rearrangements. These reactions were mainly carried out by species of clostridia and indeed some of these degradations belong to the classical age of microbiology in Delft (Liebert 1909). The number of different clostridia investigated may be gained from the following selected examples: 1. a. The classic purine-fermenting organisms Clostridium acidurici and Cl. cylindrosporum ferment several purines including uric acid, xanthine, and guanine. b. The degradation of pyrimidines such as orotic acid is accomplished by Cl. oroticum. 2. A large number of clostridia including Cl. perfringens, Cl. saccharobutyricum, Cl. propionicum, Cl. tetani, Cl. sporogenes, and Cl. tetanomorphum ferment a range of single amino acids, while some participate in the Stickland reaction involving two amino acids. 3. Several clostridia including Cl. tyrobutyricum, Cl. thermoaceticum, and Cl. kluyveri are able to reduce the double bond of α,β-unsaturated aldehydes, ketones, and coenzyme A thioesters (Rohdish et al. 2001), and are noted in Chapter 3, Part 3. Anaerobic Sulfate-Reducing Bacteria The spectrum of compounds degraded by anaerobic sulfate-reducing bacteria is continuously widening. It now includes, for example, alkanes (Aeckersberg et al. 1991, 1998), alkanols, alkylamines, and alkanoic acids, nicotinic acid (Imhoff-Stuckle and Pfennig 1983), indoles (Bak and Widdel 1986), methoxybenzoates (Tasaki et al. 1991), benzoate, hydroxybenzoate and phenol (Drzyzga et al. 1993), and catechol (Szewzyk and Pfennig 1987). Further details of the pathways are provided in Chapter 7, Part 1 and Chapter 8, Part 3. It should also be noted that elementary sulfur (S0) may serve as oxidant for organisms belonging to a number of genera including Desulfomicrobium, Desulfurella, Desulfuromonas, and Desulfuromusa (references in Liesack and Finster 1994). The anaerobic oxidation of propionate may be accomplished in pure cultures of Syntrophobacter wolinii and S. pfennigi at the expense of sulfate reduction (Wallrabenstein et al. 1995). Other Anaerobic Bacteria It is appropriate to mention briefly some additional reactions that are brought about by anaerobic bacteria: 1. Compounds such as oxalate (Dehning and Schink 1989b) and malonate (Dehning et al. 1989; Janssen and Harfoot 1992) are degraded by decarboxylation with only a modest energy contribution. They are nevertheless able to support the growth of the appropriate organisms. The bioenergetics of the anaerobic Oxalobacter formigenes (Anantharam et al. 1989) and the microaerophilic Malonomonas rubra have been elucidated (Hilbi et al. 1993; Dimroth and Hilbi 1997). 2. Particularly extensive effort has been directed to the anaerobic dechlorination of alkenes and halogenated phenols that are discussed in Chapter 7, Part 3 and Chapter 9, Part 2. As a result of this, the range of organisms that are able to carry this out is considerable and includes a. A sulfate-reducing organism Desulfomonile tiedjei (DeWeerd et al. 1990). b. A sulfite-reducing organism Desulfitobacterium dehalogenans (Utkin et al. 1994).

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c. Desulfitobacterium metallireducens that can use tetrachloroethene and trichlororethene as electron acceptors during growth with lactate (Finneran et al. 2002). d. The chlororespiring Anaeromyxobacter dehalogens that can use a few 2-chlorinated phenols as electron acceptors (Sanford et al. 2002), and in addition Fe(III) (He and Sanford 2003). e. Sulfurospirillum (Dehalospirillum) multivorans that accomplishes the sequential reduction of tetrachloroethene to trichloroethene and cis-1,2-dichloroethene (Neumann et al. 1994). 3. Although they have not been obtained in pure culture, chemolithotrophic anaerobic bacteria (anammox) that oxidize ammonia using nitrite as electron donor and CO2 as a source of carbon have been described. In addition, they can oxidize propionate to CO2 (Güven et al. 2005) by a pathway that has not yet been resolved. 4. Important groups of organisms belonging to the genus Geobacter are widely distributed in sediments. They are able to couple the reduction of Fe(III) to the oxidation of low molecularweight organic carboxylic acids (Coates et al. 1996), and some are able to use the humic acid model anthraquinone 2,6-disulfonate as electron acceptor. In addition to acetate and benzoate, G. metallireducens and G. grbiciae are able to use toluene as an electron donor with anthraquinone 2,6-disulfonate as an electron acceptor (Coates et al. 2001). These organisms are of considerable ecological importance since some of them are able to use humic acid as an electron acceptor. Further comments are given in Chapter 3, Part 3.

ORGANISMS FROM EXTREME ENVIRONMENTS Increasing numbers of bacteria have been isolated that are able to grow not only under high temperatures but also extremes of pH and salinity. There has been considerable interest in thermophiles as sources of enzymes, and, with increasing accessibility, studies in microorganisms from the high arctic from which significant features have emerged. Psychrophiles 1. Soil samples from a putatively contaminated site on Ellesmere Island (82°N, 62°W) contained bacteria belonging to the genera Pseudomonas and Sphingomonas that were psychrotolerant. They belonged to two mutually exclusive groups, one able to grow with abietanes and n-dodecane and the other with pimaranes, benzoate, and toluene. Bacteria that are able to degrade resin acids were not found in pristine tundra soils, and the isolation of the resin acid–degrading bacteria presented a conundrum (Yu et al. 2000). 2. Bacteria in thawed ice samples from Ellesmere Island (79°38′N, 74°23′W) produced CO2 and CH4 during extended incubation at 4°C in a diluted complex medium (yeast extract, casamino acids, starch, and glucose) (Skidmore et al. 2000). It was suggested that the results demonstrated that the subglacial environment beneath a polythermal glacier provided an acceptable habitat for microbial life. 3. Two different investigations have been directed to a glacial ice core (GISP2) from Greenland: a. Anomalous concentrations of methane at 2954 and 3036 m were associated with the activity of methanogens which were characterized by the fluorescence of the F420 coenzyme (Tung et al. 2005). b. Thawed samples were treated by selective filtration for ultrasmall microorganisms, followed by plating and anaerobic incubation, or by cycles of filtration and enrichment for 2–8 months at –2 or +5°C in media containing low concentrations of formate or acetate. Isolates were characterized by 16S rRNA sequencing (Miteva and Brenchley 2005). A number of important conclusions could be drawn: (i) filtration increased the number of colonies that were obtained; (ii) the combination of filtration and cultivation increased

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both the number of colonies and their diversity; and (iii) spore-forming organisms were among the isolates. Thermophiles There has been increased interest in hyperthermophiles for application to biotechnology and bioremediation. The hyperthermophilic archaeon Pyrococcus furiosus is an anaerobe that grows at temperatures up to 105°C and has a requirement for tungsten that is discussed in Chapter 3, Part 5. This organism contains several tungsten-containing ferredoxin oxidoreductases—for aldehydes, glyceraldehyde-3-phosphate, and formaldehyde (Mukund and Adams 1991, 1993, 1995; Roy et al. 1999). In addition, glutamate dehydrogenase from Pyr. furiosus is inactive below 40°C, and is able to use both NAD+ and NADP + as cofactors (Klump et al. 1992). Fe(III) reductases have been described from Pyrobaculum islandicum (Childers and Lovley 2001), and from the hyperthermophilic sulfate-reducing Archaeoglobus fulgidus (Vadas et al. 1999). Illustrative examples of their biotechnological applications using metagenomes include the following: 1. A gene (erstE1) encoding a thermostable esterase was isolated from Escherichia coli cells that had been transformed by DNA libraries with metagenomes from environmental samples isolated from thermal habitats. The enzyme belonged to the hormone-sensitive lipase family, could be overexpressed in E. coli, was active between 30 and 95°C, and used 4-nitrophenyl esters with chain lengths of C4 –C16 (Rhee et al. 2005). 2. A metagenome was isolated from soil samples, a library was constructed, and was screened for amylase activity. The amlyase gene (amyM) was overexpressed and purified. Its properties suggested that it could be regarded as a type of maltogenic amylase, α-amylase, and 4-α-glucanotransferase (Yun et al. 2004). Alkaliphiles There has been increasing interest in these on account of their potential for the synthesis of valuable enzymes that include alkaline proteases and those that degrade starch and cellulose. Strains of Bacillus sp. have been most extensively examined (Horikoshi 1999). Halophiles The organisms in a high-salinity evaporation pond were examined by plating on solid media and prolonged incubation, and by analysis of a library of PCR-amplified 16S rRNA genes (Burns et al. 2004). The isolates were related to species of several genera including Haloferax, Halorubrum, and Natromonas. The first of these was not, however, represented in the gene library, and it was suggested that this could be due to its ability to form colonies even though it was not a dominant group. A major group identified in the library was the SHOW square organisms that have so far defied cultivation.

EUKARYOTIC MICROORGANISMS: FUNGI AND YEASTS Metabolism by Fungi Although the cardinal importance of fungi in the terrestrial environment is unquestioned, and some examples of the synthesis of organohalogen metabolites are given in Chapter 3, Part 1, most attention in the aquatic environment has been directed to bacteria. An important exception is provided by studies in Lake Bonney, South Australia. The microflora of the lake contained a population of fungi including Trichoderma herzianum that was able to degrade the “free” but not the associated chloroguaiacols (van Leeuwen et al. 1997). In addition, Epicoccum sp., Mucor circinelloides, and

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Penicillium expansum—which are widely distributed soil fungi and may have entered the lake from run-off—were capable of bringing about association of tetrachloroguaiacol with organic components in the aqueous phase, so that this material could subsequently enter the sediment phase (van Leeuwen et al. 1997). On account of the similarity of the metabolic systems of fungi to those of mammalian systems, it has been suggested that fungi could be used as models for screening purposes. Several relevant examples have been provided (Ferris et al. 1976; Smith and Rosazza 1983; Griffiths et al. 1992), and the fungus Cunninghamella elegans has attracted particular attention. The reactions involved in the transformation of alachlor by this organism (Pothuluri et al. 1993) are similar to those encountered in other organisms: (a) mammalian systems, (b) the biotransformation of the analogous metalochlor by bluegill sunfish (Lepomis macrochirus), and (c) that carried out by a soil actinomycete, which is noted later in this chapter. A wide range of taxonomically diverse fungi has also been used for the synthesis of less readily available compounds including hydroxylated steroids (Chapter 7, Part 2). Fungi—especially white-rot and brown-rot —have attracted considerable attention in the context of bioremediation of contaminated terrestrial sites, and examples are given in Chapter 14. In the terrestrial environment—and possibly also in a few specialized aquatic ecosystems— fungi and yeasts play a cardinal role in biodegradation and biotransformation. The role of yeasts in the coastal marine environment is illustrated by results of their frequency and their potential for transformation of phenanthrene and benz[a]anthracene (MacGillivray and Shiaris 1993). The transformation of PAHs by certain fungi is analogous to that in mammalian systems so that fungal metabolism has been explored as a model for higher organisms, and extensive studies have been carried out, particularly with Cunninghamella elegans. It has been suggested that the transformations accomplished by the brown-rot fungus Gleophyllum striatum may involve hydroxyl radicals, and this is supported by the overall similarity in the structures of the fungal metabolites with those produced with Fenton’s reagent (Wetzstein et al. 1997). Biotransformation (hydroxylation) of a wide range of PAHs and related compounds including biphenyl, naphthalene, anthracene, phenanthrene, 4-, and 7-methylbenz[a]anthracene, and 7, 12-dimethylbenz[a]anthracene has been examined in a number of fungi, most extensively in species of Cunninghamella, especially C. elegans (McMillan et al. 1987). Hydroxylation of benzimidazole (Seigle-Murandi et al. 1986) by Absidia spinosa, and of biphenyl ether by C. echinulata (SeigleMurandi et al. 1991) has also been studied on account of the industrial interest in the metabolites. The biotransformation of alachlor (2-chloro-N-methoxymethyl-N-[2,6-diethylphenyl]acetamide) by C. elegans has already been mentioned, and involves primarily hydroxylation at the benzylic carbon atom and loss of the methoxymethyl group (Pothuluri et al. 1993). In all these examples, the reactions are correctly described as biotransformations since the aromatic rings of these compounds remain unaltered. Particular attention has been focused on the white-rot fungus Phanerochaete chrysosporium on account of its ability to degrade not only lignin but also to metabolize a wide range of unrelated compounds. These include PAHs (Bumpus 1989) and organochlorine compounds including DDT (Bumpus and Aust 1987), PCBs (Eaton 1985), lindane and chlordane (Kennedy et al. 1990), pentachlorophenol (Mileski et al. 1988), and 2,7-dichlorobenzo[1,4]dioxin (Valli et al. 1992b). The novel pathways for the degradation of 2,4-dichlorophenol (Valli and Gold 1991) (Chapter 9, Part 2) and 2,4-dinitrotoluene (Valli et al. 1992a) (Chapter 9, Part 5) are worth noting. The degradation of all these compounds is apparently mediated by two peroxidase systems—lignin peroxidases and manganese-dependent peroxidases—and a laccase that is produced by several white-rot fungi, though not by P. chrysosporium. A potentially serious interpretative ambiguity has, however, emerged from the observation that lignin peroxidase is able to polymerize a range of putative aromatic precursors to lignin, but that this is not the functional enzyme in the depolymerization of lignin (Sarkanen et al. 1991). The regulation of the synthesis of these oxidative enzymes is complex, and is influenced by nitrogen limitation, the growth status of the cells, and the concentration of manganese in the medium (Perez and Jeffries 1990). In addition, it seems clear that monooxygenase and epoxide

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hydrolase activities are also involved since the biotransformation of phenanthrene takes place even in the absence of peroxidase systems (Sutherland et al. 1991). P. chrysosporium is able to degrade anthracene by oxidation to anthra-9,10-quinone, followed by ring fission to produce o-phthalate. Phanerochaete chrysosporium, which is able to degrade simultaneously chlorobenzene and toluene (Yadav et al. 1995) has achieved some importance in the context of bioremediation since the bacterial degradation of these substrates is generally restricted by the incompatibility of the degradative pathways. Clearly therefore, a number of important unresolved issues remain. In addition, attention is drawn to the role of fungal redox systems in the covalent linking of xenobiotics to aromatic components of humus in soils. The metabolic activity of other white-rot fungi including Phanerochaete chrysosporium and Pleurotus ostreacus has been discussed in the context of polycyclic aromatic hydrocarbons. For example, the mineralization potential of the manganese peroxide system from Nematoloma frowardii for a number of substrates has been demonstrated (Hofrichter et al. 1998): the formation of CO2 from labeled substrates ranged from 7% (pyrene) to 36% (pentachlorophenol), 42% (2-amino-4, 6-dinitrotoluene), and 49% (catechol). Several strains of white-rot fungi have been examined for their ability to degrade and mineralize selected PCB congeners (Beaudette et al. 1998). Mineralization of 2,4′,5 [U-14C]trichlorobiphenyl as a fraction of the substrate added ranged from ca. 4.2% for a strain of Pleurotus ostreacus, 4.9 and 6.9% for two strains of Bjerkandera adusta, to 11% for a strain of Trametes versicolor, whereas Phanerochaete chrysosporium produced only ca. 2% of 14CO2. There was no apparent correlation among levels of lignin peroxidase, manganese peroxidase, and degradative ability. The degradation of phenolic compounds by fungi may involve rather unusual features of which the following three are given as illustration: 1. Aspergillus fumigatus degrades phenol using simultaneously two pathways: first, orthohydroxylation to catechol followed by ring cleavage to 3-ketoadipate, and second, successive hydroxylation to hydroquinone and 1,2,4-trihydroxybenzene before ring cleavage (Jones et al. 1995). The metabolism of 2-aminobenzoate (anthranilate) to 2,3-dihydroxybenzoate by A. niger is accomplished by apparent incorporation of one atom of oxygen from each of O2 and H2O, and is not a flavoprotein (Subramanian and Vaidyanathan 1984). 2. The degradation of 1,3,5-trihydroxybenzene (phloroglucinol) by Fusarium solani involves rearrangement to 1,2,3-trihydroxybenzene (pyrogallol) followed by ring cleavage to 2-ketohex-3-ene-1,6-dicarboxylate (Walker and Taylor 1983). This rearrangement is the opposite of that involved in the anaerobic degradation of 3,4,5-trihydroxybenzoate (Chapter 8, Part 4). 3. The imperfect fungus, Paecilomyces liliacinus, is able to produce mono-, di-, and trihydroxylated metabolites successively from biphenyl, and carry out intradiol ring fission of rings carrying adjacent hydroxyl groups (Gesell et al. 2001). Similarly, mono- and dihydroxylated metabolites are produced from dibenzofuran followed by intradiol fission of the ring with adjacent hydroxyl groups (Gesell et al. 2004). Metabolism by Yeasts The metabolic potential of yeasts has attracted attention in different contexts, and it has emerged that in contrast to many fungi, they are able to bring about fission of aromatic rings. Some examples that illustrate the various possibilities are given below: 1. Ring fission clearly occurs during the metabolism of phenol (Walker 1973) by the yeast Rhodotorula glutinis, and of aromatic acids by various fungi (Cain et al. 1968; Durham et al. 1984; Gupta et al. 1986). 2. Analogous ring fission reactions have also been found in studies on the metabolism of other aromatic compounds by the yeast Trichosporon cutaneum whose metabolic versatility is indeed comparable with that of bacteria. Examples include the degradation of

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phenol and resorcinol (Gaal and Neujahr 1979), tryptophan and anthranilate (Anderson and Dagley 1981a), and aromatic acids (Anderson and Dagley 1981b). 3. The ability to grow at the expense of 4-hydroxy- and 3,4-dihydroxybenzoate has been used for the classification of medically important yeasts including Candida parapsilosis (Cooper and Land 1979). This organism degrades these substrates by oxidative decarboxylation, catalyzed by a flavoprotein monooxygenase (Eppink et al. 1997). 4. The phenol-assimilating yeast Candida maltosa degraded a number of phenols, even though these were unable to support growth. Hydroxylation of 3-chloro- and 4-chlorophenol initially produced 4-chlorocatechol and then 5-chloropyrogallol (Polnisch et al. 1992). The yeast Rhodotorula cresolica was able to assimilate phenol, 3- and 4-methylphenol, catechol and 3- and 4-methylcatechol, resorcinol and hydroquinone, and a wide range of phenolic carboxylic acids (Middelhoven and Spaaij 1997). 5. Diphenyl ether is transformed by Trichosporum beigelii through successive hydroxylation and extradiol ring fission of the resulting catechol (Schauer et al. 1995). In a formally analogous way, the metabolism of dibenzofuran by the yeast Trichosporon mucoides involves initial hydroxylation of one of the rings followed by ring fission at the 2,3-position (Hammer et al. 1998). Yeasts are also able to degrade long-chain alkanes. This is accomplished in two subcellular organelles—in microsomes, cytochrome P450 and the associated NADH reductase (Käppeli 1986) carry out hydroxylation, while the alkanols are dehydrogenated and undergo β-oxidation in peroxisomes that are induced during growth with alkanes (Tanaka and Ueda 1993). Clearly then, yeasts possess metabolic potential for the degradation of xenobiotics little inferior to that of many bacterial groups, so that their role in natural ecosystems justifies the greater attention that has been directed to them (MacGillivray and Shiaris 1993).

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Vannelli T, AB Hooper (1992) Oxidation of nitrapyrin to 6-chloropicolinic acid by the ammonia-oxidizing bacterium Nitrosomonas europaea. Appl Environ Microbiol 58: 2321–2325. Vannelli T, M Logan, DM Arciero, AB Hooper (1990) Degradation of halogenated aliphatic compounds by the ammonia-oxidizing bacterium Nitrosomonas europaea. Appl Environ Microbiol 56: 1169–1171. Vercellone-Smith P, DS Herson (1997) Toluene elicits a carbon starvation response in Pseudomonas putida mt-2 containing the TOL plasmid pWWO. Appl Environ Microbiol 63: 1925–1932. Verhagen FJM, HJ Swarts, JBPA Wijnberg, JA Field (1998) Biotransformation of the major fungal metabolite 3,5-dichloro-p-anisyl alcohol under anaerobic conditions and its role in formation of bis(3,5-dichloro4-hydroxyphenyl)methane. Appl Environ Microbiol 64: 3225–3231. Véron M, L Le Minor (1975) Nutrition et taxonomie des Enterobacteriaceae et bactéries voisines III Caractéres nutritionnels et différenciation des groupes taxonomiques. Ann Microbiol (Inst Pasteur) 126B: 125–147. Wackett LP, GA Brusseau, SR Householder, RS Hanson (1989) Survey of microbial oxygenases: trichloroethylene degradation by propane-oxidizing bacteria. Appl Environ Microbiol 55: 2960–2964. Wahman DG, LE Katz, GE Speitel (2005) Cometabolism of trihalomethanes by Nitrosomonas europaea. Appl Environ Microbiol 71: 7980–7986. Walker JRL, BG Taylor (1983) Metabolism of phloroglucinol by Fusarium solani. Arch Microbiol 134: 123–126. Walker N (1973) Metabolism of chlorophenols by Rhodotorula glutinis. Soil Biol Biochem 5: 525–530. Wallrabenstein C, E Hauschild, B Schink (1995) Syntrophobacter pfennigii sp. nov., new syntrophically propionate-oxidizing anaerobe growing in pure culture with propionate and sulfate. Arch Microbiol 164: 346–352. Wang Y, PCK Lau, DK Button (1996) A marine oligobacterium harboring genes known to be part of aromatic hydrocarbon degradation pathways of soil pseudomonads. Appl Environ Microbiol 62: 2169–2173. Ward BB (1987) Kinetic studies on ammonia and methane oxidation by Nitrosococcus oceanus. Arch Microbiol 147: 126–133. Warshawsky D, M Radike, K Jayasimhulu, T Cody (1988) Metabolism of benzo(a)pyrene by a dioxygenase system of the freshwater green alga Selenastrum capricornutum. Biochem Biophys Res Comm 152: 540–544. Warshawsky D, T Cody, M Radike, R Reilman, B Schumann, K LaDow, J Schneider (1995) Biotransformation of benzo[a]pyrene and other polycyclic aromatic hydrocarbons and heterocyclic analogues by several green algae and other algal species under gold and white light. Chem-Biol Interact 97: 131–148. Wayne LG et al. (1991) Fourth report of the cooperative, open-ended study of slowly growing mycobacteria by the international working group on mycobacterial taxonomy. Int J Syst Bacteriol 41: 463–472. Wedemeyer G (1967) Dechlorination of 1,1,1-trichloro-2,2-bis[p-chlorophenyl]ethane by Aerobacter aerogenes. Appl Microbiol 15: 569–574. Wetzstein H-G, N Schmeer, W Karl (1997) Degradation of the fluoroquinolone enrofloxacin by the brown-rot fungus Gleophyllum striatum: identification of metabolites. Appl Environ Microbiol 63: 4272–4281. Whittenbury R, DP Kelly (1977) Autotrophy: a conceptual phoenix. Symp Soc Gen Microbiol 27: 121–149. Wiberg K, E Brorstöm-Lundén, TF Bidleman, P Haglund (2001) Concentrations and fluxes of hexachlorocyclohexanes and chiral composition of α-HCH in environmental samples from the southern Baltic Sea. Environ Sci Technol 35: 4739–4746. Wieser M, B Wagner, J Eberspächer, F Lingens (1997) Purification and characterization of 2,4,6-trichlorophenol-4-monooxygenase, a dehalogenating enzyme from Azotobacter sp. strain GP1. J Bacteriol 179: 202–208. Wong CM, MJ Dilworth, AR Glenn (1994) Cloning and sequencing show that 4-hydroxybenzoate hydroxylase Poba is required for uptake of 4-hydroxybenzoate in Rhizobium leguminosarum. Microbiology (UK) 140: 2775–2786. Yadav JS, RE Wallace, CA Reddy (1995) Mineralization of mono- and dichlorobenzenes and simultaneous degradation of chloro- and methyl-substituted benzenes by the white-rot fungus Phanerochaete chrysosporium. Appl Environ Microbiol 61: 677–680. Yakimov MM, PN Golyshin, S Lang, ERB Moore, W-R Abraham, H Lünsdork, KN Timmis (1998) Alcanivorax borkumensis gen. nov., sp. nov., a new hydrocarbon-degrading and surfactant-producing marine bacterium. Int J Syst Bacteriol 48: 339–348. Yoshinaga DH, HA Frank (1982) Histamine-producing bacteria in decomposing skipjack tuna (Katsuwonus pelamis). Appl Environ Microbiol 44: 447–452. Yoshioka T, JA Krauser, FP Guengerich (2002) Tetrachloroethylene oxide: hydrolytic products and reactions with phosphate and lysine. Chem Res Toxicol 15: 1096–1105.

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Yu Z, GR Stewart, W Mohn (2000) Apparent contradiction: psychrotolerant bacteria from hydrocarboncontaminated arctic tundra soils that degrade diterpenoids synthesized by trees. Appl Environ Microbiol 66: 5148–5154. Yun J, S Kang, S Park, H Yoon, M-J Kim S Heu, S Ryu (2004) Characterization of a novel amylolytic enzyme encoded by a gene from a soil-derived metagenomic library. Appl Environ Microbiol 70: 7229–7234. Zaborina O, M Latus, J Eberspächer, LA Golovleva, F Lingens (1995) Purification and characterization of 6-chlorohydroquinol 1,2-dioxygenase from Streptomyces rochei 303: comparison with an analogous enzyme from Azotobacter sp. strain GP1. J Bacteriol 177: 229–234. Zaitsev GM, JS Uotila, IV Tsitko, AG Lobanok, MS Salkinoja-Salonen (1995) Utilization of halogenated benzenes, phenols, and benzoates by Rhodococcus opacus GM-14. Appl Environ Microbiol 61: 4191–4201. Zhao J-S, A Halasz, L Paquet, C Beaulieu, J Hawari (2002) Biodegradation of hexahydro-1,3,5-trinitro-1,3,5triazine and its mononitroso derivative hexahydro-1-nitroso-3,5-dinitro-1,3,5-triazine by Klebsiella pneumoniae strain SCZ isolated from an anaerobic sludge. Appl Environ Microbiol 68: 5336–5341. Zikmundová M, K Drandarov, L Bigler, M Hesse, C Werner (2002) Biotransformation of 2-benzoxazolinone and 2-hydroxy-1,4-benzoxazin-3-one by endophytic fungi isolated from Aphelandra tetragona. Appl Environ Microbiol 68: 4863–4870.

PART 2 REACTIONS MEDIATED BY HIGHER ORGANISMS INTRODUCTION Very few xenobiotics remain unaltered in the environment for any length of time after their release. Although metabolism by microorganisms has already been discussed, brief comments are made here on the metabolism of contaminants by higher organisms—particularly fish. Metabolites such as conjugates may enter the aquatic environment where their ultimate fate is determined by microbial reactions. Metabolism by higher organisms is relevant to the mechanisms whereby organisms detoxify deleterious contaminants or, conversely, induce the synthesis of toxic metabolites. Some illustrations of these are briefly summarized below: 1. The kinetics and products of metabolism critically influence the nature and the concentrations of the xenobiotic and its transformation products to which the cells are exposed. Increasing evidence from different sources has shown that the effective toxicant may indeed be a metabolite synthesized from the compound originally supplied and not the xenobiotic itself. It is important to appreciate that, as with toxicity, the extent of metabolism will generally depend on the nature and position of substituents on aromatic rings as well as on their number. For example, although 2,3,4- and 3,4,5-trichloroaniline were N-acetylated in guppy, this did not occur with 2,4,5-trichloroaniline (de Wolf et al. 1993). Metabolism may also be an important determinant of genotoxic effects, estrogenic activity, and teratogenicity. 2. The classic example is that of Prontosil (Figure 2.12) in which the compound is active against bacterial infection in animals though inactive against the bacteria in pure culture. The toxicity in animals is the result of reduction to the sulfanilamide (4-aminobenzenesulfonamide) that competitively blocks the incorporation of 4-aminobenzoate into the vitamin folic acid. 3. The mechanism of fluoroacetate toxicity in mammals has been extensively examined and was originally thought to involve simply initial synthesis of fluorocitrate that inhibits aconitase and thereby the functioning of the TCA cycle (Peters 1952). Walsh (1982) has

NH 2

NH 2 H2 N

FIGURE 2.12

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N

N

SO 2 NH 2

H2 N

NH 2 + H 2 N

SO 2 NH 2

Metabolism of Prontosil in mammals.

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FIGURE 2.13

4.

5.

6.

7.

C

S − CH 3 CH

N −O −CONHCH3

CH 3 CH 3

C

S − CH 3 CH

N−O−CONHCH3

Metabolism of aldicarb by rainbow trout.

extensively reinvestigated the problem, and revealed both the complexity of the mechanism of inhibition and the stereospecificity of the formation of fluorocitrate from fluoroacetate. Considerable attention has been directed to the synthesis of the epoxides and dihydrodiol epoxides of polycyclic aromatic hydrocarbons mediated by the action of cytochrome P450 systems, and their role in inducing carcinogenesis in fish (Varanasi et al. 1987; de Maagd and Vethaak 1998). Tumors observed in feral fish exposed to PAHs may plausibly—though not necessarily—be the result of this transformation. Even though an apparently causal relationship between exposure of fish to PAHs and disease may have been established (Malins et al. 1985, 1987), caution should be exercised due to the possibility that other— and unknown—substances may have induced carcinogenesis. It is also important to appreciate that other compounds may induce induction of the metabolic system for PAHs. For example, exposure of rainbow trout to PCBs increases the effectiveness of liver enzymes to transform benzo[a]pyrene to carcinogenic intermediates (Egaas and Varanasi 1982). The carbamate insecticide aldicarb (Figure 2.13) that exerts its effect by inactivating acetylcholinesterase is metabolized by a flavin monooxygenase from rainbow trout to the sulfoxide, which is a more effective inhibitor (Schlenk and Buhler 1991). Pre-exposure to the organophosphate diazinon at exposures half the LC50 values increased the LC50 value by a factor of about five for guppy (Poecilia reticulata), but had no effect on the value for zebra fish (Brachydanio rerio). This was consistent with the observation that during pre-exposure of guppy there was a marked inhibition in the synthesis of the toxic metabolites diazoxon and pyrimidinol, whereas this did not occur with zebra fish in which the toxicity was mediated primarily by the parent compound (Keizer et al. 1993). Pyrene is metabolized by the fungus Crinipellis stipitaria to 1-hydroxypyrene, and this has a spectrum of toxic effects substantially greater than those of pyrene: these include cytotoxicity to HeLa S3 cells, toxicity to a number of bacteria and to the nematode Caenorhabditis elegans (Lambert et al. 1995).

At the other extreme, if metabolism of the xenobiotic by the organism does not occur at all—or at insignificant rates—after exposure, the compound will be persistent in the organism, and may therefore be consumed by predators. This is relevant to biomagnification. Most of the reactions carried out by fish and higher aquatic organisms are relatively limited transformation reactions in which the skeletal structure of the contaminants remains intact. The following three widely distributed reactions are of greatest significance: 1. Cytochrome P450–type monooxygenase systems, which have a generally low substrate specificity, are widely distributed in the species of fish used for toxicity testing (Funari et al. 1987). 2. Glutathione S-transferases (Donnarumma et al. 1988; Nimmo 1987), which are important in the metabolism of highly reactive compounds containing electrophilic groups such as epoxides, and aromatic rings with several strongly electron-attracting substituents such as halogen, cyano, or nitro groups. 3. Conjugation of polar groups such as amines, carboxylic acids, and phenolic hydroxyl groups produce water-soluble compounds that are excreted and these reactions therefore function as a detoxification mechanism.

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METABOLISM BY FISH The biotransformation of xenobiotics in many higher organisms is mediated by the cytochrome P450 monooxygenase system and the complexity of factors that regulate the synthesis of this in fish has been reviewed (Andersson and Förlin 1992). The metabolic potential of fish may appear restricted compared with that of microorganisms, but it may have been considerably underestimated. For example, metolachlor (2-chloro-N-[2-ethyl-6-methylphenyl]-N-[2-methoxy-1methylethyl]acetamide) is metabolized by bluegill sunfish (Lepomis macrochirus) by reactions involving initially O-demethylation and hydroxylation (Cruz et al. 1993) (Figure 2.14). These are comparable to the reactions carried out by an actinomycete (Krause et al. 1985), and the benzylic hydroxylation is analogous to that involved in the biotransformation of the structurally similar alachlor by the fungus Cunninghamella elegans (Pothuluri et al. 1993). An extensive compilation of the transformation of xenobiotics by fish has been given (Sijm and Opperhuizen 1989), and only a few examples of these reactions are summarized here as illustration: 1. N-Dealkylation of dinitramine to 1,3-diamino-2,4-dinitro-6-trifluoromethylbenzene (Olson et al. 1977) by carp (Cyprinus carpio) (Figure 2.15). 2. O-Demethylation of pentachloroanisole in rainbow trout (Salmo gairdneri ≡ Oncorhynchus mykiss) (Glickman et al. 1977), and of chlorinated veratroles by zebra fish (Neilson et al. 1989). 3. Acetylation of 3-amino ethylbenzoate in rainbow trout (Hunn et al. 1968). 4. Displacement of the nitro group in pentachloronitrobenzene by hydroxyl and thiol groups (Bahig et al. 1981) (Figure 2.16) in golden orfs (Idus idus). 5. Oxidation of a number of PAHs has been demonstrated in a variety of fish. A review directed to metabolism and the role of PAH metabolites in inducing tumorogenesis has been given in de Maagd and Vethaak (1998), and only two examples are given here: a. Coho salmon (Oncorhynchus kisutch) metabolized naphthalene to a number of compounds consistent with oxidation to the epoxide, hydrolysis to the dihydrodiol, and dehydration of the trans-dihydrodiol to naphth-1-ol, or by rearrangement of the epoxide (Figure 2.17) (Collier et al. 1978). b. For the carcinogen benzo[a]pyrene, a wider range of metabolites has been identified in southern flounder (Paralichthys lethostigma) including the 4,5-, 7,8-, and 9,10diols, the 1,6-, 3,6-, and 6,12-quinones in addition to the 1-, 3-, and 9-benzopyreneols (Figure 2.18) (Little et al. 1984). CH 3

CH 3

CH 3 CH 3

CH 2

CH−CH2−O−CH3 N

N

COCH 2 Cl

CH−CH2OH

CH−CH2OH COCH 2 Cl

CH 2 OH

Metabolism of metolachlor by bluegill sunfish.

F3 C

NH 2 NO 2

F3C

NO 2

NH2 NO 2

N(C2 H 5 )2

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N

CH 3

NH 2

FIGURE 2.15

CH 3

COCH 2 Cl

CH 3

FIGURE 2.14

CH 2

CH 3

CH 2

F3 C

NO2

NHEt NO 2

NH2 NO2

Metabolism of dinitramine by carp.

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Cl

Cl

Cl

NO 2

+

Conjugate

+

Conjugate

Cl

Cl

Cl

Cl

Cl

SH

Cl

Cl

Cl

Cl

Cl Cl

FIGURE 2.16

Metabolism of pentachloronitrobenzene by golden orfs. H

OH H

O H

FIGURE 2.17

OH

OH H

Metabolism of naphthalene by coho salmon. OH OH

OH OH

OH OH

O

O

O

O

FIGURE 2.18

O

O

Metabolism of benzo[a]pyrene by southern flounder.

6. N-Hydroxylation of aniline and 4-chloroaniline by rainbow trout to hydroxylamines that could plausibly account for the subchronic toxicity of the original compounds (Dady et al. 1991). Initially formed polar metabolites such as phenols and amines may be conjugated to water-soluble terminal metabolites that are excreted into the medium and function as an effective mechanism of detoxification. For example, pentachlorophenol and pentachlorothiophenol produced from pentachloronitrobenzene conjugated represented the major metabolites. Although the naphthalene dihydrodiol was the major metabolite produced from naphthalene, the further transformation

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O−CH2−CO−NH−CH2−CH2−SO3H

Cl

Cl

FIGURE 2.19

Cl

Cl

Conjugation of carboxylic acids with amino acids.

product naphth-1-ol was also isolated as the sulfate, glucuronate, and glucose conjugates. Diverse conjugation reactions have been described and include the following: 1. Phenolic compounds with formation of glucuronides, sulfates, or glucosides as already noted. 2. Reaction of carboxylic acids with the amino groups of glycine (Huang and Collins 1962) or taurine (Figure 2.19) (James and Bend 1976) to form the amides. 3. Reaction between glutathione and reactive chloro compounds such as 1-chloro-2,4-dinitrobenzene (Niimi et al. 1989), or the chloroacetamide group in demethylated metolachlor (Cruz et al. 1993). Important investigations have been directed to persistent halogenated aromatic compounds and different mechanisms for their metabolism have been found: 1. Hydroxylated PCBs have been found in a laboratory study using rainbow trout (Oncorhynchus mykiss), and were similar to those previously observed (Campbell et al. 2003) in wild lake trout (Salvelinus namaycush). Although it was not possible to associate the metabolites unambiguously with their precursor PCB congener, those with neighboring hydrogen atoms in the ortho/para positions were most probably transformed, putatively by CYP 2B-type isoforms of cytochrome P450 rather than by CYP 1A enzymes (Buckman et al. 2006). 2. Debromination has been examined in rainbow trout (Oncorhynchus mykiss) and carp (Cyrprinus carpio) that were fed with a diet containing decabromodiphenyl ether (BDE 209). In both of them, products from debromination were found in whole fish, whole fish homogenates, organs, and microsomal preparations. In trout, debromination produced primarily the octa and nano congeners, whereas in carp penta to octa congeners were found. This was confirmed in microsomal preparations of carp that transformed up to 63% of DBE 209 with the formation of hexa to octa congeners, whereas in trout only 23% of DBE 209 was transformed to octa and nano congeners (Stapleton et al. 2006). It was suggested that the debromination could be analogous to the deiodination of thyroxine to the more active triiodothyronine by loss of iodine from the position ortho to the hydroxyl group.

METABOLISM BY OTHER ORGANISMS Mussels 1. Mussels do not generally carry out more than the limited reactions of oxidation and conjugation, and in the common mussel, Mytilus edulis, variations between summer and winter levels for both cytochrome P450 and NADPH-independent 7-ethoxycoumarin O-deethylase have been found (Kirchin et al. 1992). Levels of cytochrome P450 and the rates of metabolism of PAHs were apparently low compared with those found in fish (Livingstone and Farrar 1984), and an investigation using subcellular extracts of the digestive glands

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from the mussel M. galloprovincialis showed that although the formation of diols and phenols from benzo[a]pyrene was dependent on NADPH, the quinones that were the major metabolites were produced in the absence of NADPH apparently by radical-mediated reactions involving lipid peroxidase systems (Michel et al. 1992). 2. A reaction presumably mediated by glutathione S-transferase is the replacement of the 4-chloro substitutent in octachlorostyrene in the blue mussel (Mytilus edulis) by a thiomethyl group (Figure 2.20) (Bauer et al. 1989). A similar reaction of glutathione with arene oxides produced by aquatic mammals from PCBs and DDE results ultimately in the production of methyl sulfones (Bergman et al. 1994; Janák et al. 1998). In the arctic food chain—arctic cod (Boreogadus saida), ringed seal (Phoca hispidus), polar bear (Ursus maritima)—it has been shown that levels of the dimethyl sulfones of DDE and PCB were low in cod, and that levels in polar bear were the combined result of bioaccumulation from seals and endogenous metabolism (Letcher et al. 1998). The results of these investigations suggest that caution should be exercised in interpreting not only the results of toxicity assays in which such organisms are employed but also data accumulated in monitoring studies that may not have taken into account the existence of metabolites.

INSECTS Resistance of house flies (Musca domestica) to DDT was attributed to its transformation to the nontoxic DDE, and the enzyme that carries the dehydrochlorination has been characterized in DDTresistant flies (Lipke and Kearns 1959a,b). The herbicide alachlor is transformed by chironomid larvae by O-demethylation followed by loss of the chloroacetyl group to produce 2,6-diethylaniline (Figure 2.21) (Wei and Vossbrinck 1992).

INVERTEBRATES The metabolism of xenobiotics by both terrestrial and sediment-dwelling biota has been studied, and provides illustrations of the importance of uptake by food or by sorbed sediment. Some examples of metabolism by terrestrial biota include the following.

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

FIGURE 2.20

Et

SCH 3

Metabolism of octachlorostyrene by the blue mussel.

CH2−O −CH3

Et

Et CH2−OH

N

CO −CH2Cl Et

Et NH −CO −CH2Cl

N CO −CH2Cl

Et

Cl

Et

NH2 Et

FIGURE 2.21 Metabolism of alachlor by chironomid larvae.

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Isopods The uptake and elimination of benzo[a]pyrene by the terrestrial isopod Porcellio scaber have been investigated (van Brummelen and van Straalen 1996), and 1-hydroxypyrene was identified among the metabolites of pyrene in this organism (Stroomberg et al. 1996). Oligochaetes Both the (+)- and (–)- enantiomers of limonene were transformed by larvae of the cutworm Spodoptera litura (Miyazawa et al. 1998). For both of them the reactions involved are (a) dihydroxylation between C-8 and C-9 and (b) oxidation of the C-1 methyl group to carboxyl. These transformations were not dependent on the intestinal microflora in contrast to the transformation of α-terpinene to p-mentha-1,3-dien-7-ol and p-cymene whose formation could be attributed to the intestinal flora. Polychaetes Polychaete worms belonging to the genera Nereis and Scolecolepides have extensive metabolic potential. Nereis virens is able to metabolize PCBs (McElroy and Means 1988) and a number of PAHs (McElroy 1990), while N. diversicolor and Scolecolepides viridis are able to metabolize benzo[a]pyrene (Driscoll and McElroy 1996). It is worth noting that apart from excretion of the toxicant, polar, and much more water-soluble metabolites such as the glycosides formed from pyrene by Porcellio sp. (Larsen et al. 1998) may be mobile in the interstitial water of the sediment phase. It has been shown that although the marine terebellid polychaete Amphitrite ornata produced no detectable halogenated metabolites, it synthesizes a dehalogenase that is able to carry out oxidative dehalogenation of halogenated phenols, with fluoro, chloro, or bromo substituents (Chen et al. 1996). One of the enzymes (DHB 1) has been purified, and consists of two identical subunits (Mr 15,530) each containing heme and histidine as the proximal Fe ligand. In the presence of H2O2, 2,4,6-tribromophenol is oxidized to 2,6-dibromo-benzo-1,4-quinone (LaCount et al. 2000).

HIGHER PLANTS Attention is briefly drawn to the metabolism of contaminants by higher plants for several reasons: a. The metabolites may be toxic to biota at higher trophic levels. b. The metabolites may be translocated into the root system and, after partition into interstitial water in the soil, exert a toxic effect on other terrestrial organisms. c. There has been interest in the use of higher plants in the bioremediation of contaminated sites. This issue is addressed in detail in Chapters 12 and 14. The detoxification and metabolism of higher plants has been reviewed (Sandermann 1994), and a few examples are given below as illustration: 1. Pentachlorophenol is metabolized by the aquatic plant Eichhornia crassipes to a number of metabolites including di-, tri-, and tetrachlorocatechol, 2,3,5-tri- and tetrachlorohydroquinone, pentachloroanisole, and tetrachloroveratrole (Roy and Hänninen 1994). The phenolic compounds should be compared with those produced during the photochemical transformation (Figure 2.22) and the initial stages in the microbiological metabolism of pentachlorophenol, followed by O-methylation. 2. Quite complex transformations may be mediated; for example, the metabolism of phoxim by plant organs and cell suspension of soybean (Glycine max) (Höhl and Barz 1995) (Figure 2.23).

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Environmental Degradation and Transformation of Organic Chemicals OH Cl

OH

Cl

Cl

O Cl

Cl O

OH Cl

Cl

Cl

Cl

Cl

Cl

Cl

HO

Cl

O Cl

Cl

CO2H

Cl

CO2H

O

Cl

OH

FIGURE 2.22

OH

Cl

Cl

Cl

Cl

Photochemical transformation of pentachlorophenol. CH 3 N

CH 3 O

OEt P

NOH

OEt

S

CH 3 NH 2

CH 3 NH−CO−CH2−CO2H

N-malonylation

FIGURE 2.23

Metabolism of phoxim by soybean (Glycine max).

3. The uptake and biotransformation of benzene from soil and the atmosphere has been studied in a number of plants. It was shown that in leaves of spinach (Spinacia oleracea) the label in 14C-benzene was found in muconic, fumaric, succinic, malic, and oxalic acids, as well as in specific amino acids, and that an enzyme preparation in the presence of NADH or NADPH produced phenol (Ugrekhelidze et al. 1997). 4. Hybrid poplars are able to transport and metabolize diverse xenobiotics: a. Trichloroethene was metabolized to trichloroethanol and trichloroacetate (Newman et al. 1997), and the trichloroethanol conjugated to the β-d-glucoside (Shang et al. 2001). b. Atrazine was metabolized by reactions involving dealkylation and hydrolytic dechlorination to 4,6-diamino-1,3,5-triazine (Burken and Schnoor 1997). c. Tissue cultures of poplar (Populus deltoides × nigra) metabolized the explosive hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) by partial reduction to the 1-nitroso and 1,3-dintroso derivatives, and in the light these were further metabolized to CH2O, CH3OH, and CO2 (van Aken et al. 2004). 5. 2,4,6-Trinitrotoluene (TNT) is reduced by the aquatic plant Myriophyllum spicatum to aminodinitrotoluenes (Pavlostathis et al. 1998) and, in axenic root cultures of Catharanthus roseus, the initial metabolites 2-amino-4,6-dinitrotoluene and 4-amino-2,6-dintrotoluene

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were conjugated probably with C6-units (Bhadra et al. 1999). There are clearly several important unresolved issues including the phytotoxicity of these metabolites before phytoremediation of TNT could be exploited. 6. Reed Canary Grass (Phalaris arundinacea) was grown in liquid culture and exposed to RDX that was metabolized to the potentially toxic 4-nitro-2,4-diazabutanal (Just and Schnoor 2004). This metabolite is also produced from RDX by strains of Rhodococcus sp. and from the homologous octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine (HMX) by Phanerochaete chrysosporium. 7. Transformation of DDT to DDD by reductive dechlorination has been demonstrated in a number of aquatic plants, although the reaction appears to be abiotic mediated by some component of the plants (Garrison et al. 2000).

OTHER ORGANISMS Other groups of biota are able to bring about transformation of structurally diverse compounds and limited investigations have revealed the metabolic potential of taxonomically diverse eukaryotic organisms: a. The apochlorotic alga (protozoan) Prototheca zopfii is able to degrade aliphatic hydrocarbons (Walker and Pore 1978; Koenig and Ward 1983). b. Tetrahymena thermophila transforms pentachloronitrobenzene to the corresponding aniline and pentachlorothioanisole (Figure 2.24) (Murphy et al. 1982). c. Daphnia magna has been shown to bring about dechlorination and limited oxidation of heptachlor (Figure 2.25) (Feroz et al. 1990).

NO 2

NH 2

SCH 3

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

+ Cl

FIGURE 2.24

Cl

Cl

Biotransformation of pentachloronitrobenzene by Tetrahymena thermophila. Cl

Cl Cl

Cl Cl

Cl Cl

Cl

Cl

Cl

Cl

O Cl Cl

Cl Cl

Cl Cl

Cl

OH

Cl

OH Cl

FIGURE 2.25

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Biotransformation of heptachlor by Daphnia magna.

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REFERENCES Andersson T, L Förlin (1992) Regulation of the cytochrome P450 enzyme system in fish. Aquat Toxicol 24: 1–20. Bahig ME, A Kraus, W Klein, F Korte (1981) Metabolism of pentachloronitrobenzene-14C quintozene) in fish. Chemosphere 10: 319–322. Bauer I, K Weber, W Ernst (1989) Metabolism of octachlorostyrene in the blue mussel (Mytilus edulis). Chemosphere 18: 1573–1579. Bergman Å, RJ Norstrom, K Haraguchi, H Kuroki, P Béland (1994) PCB and DDE methyl sulfones in mammals from Canada and Sweden. Environ Toxicol Chem 13: 121–128. Bhadra R, DG Wayment, JB Hughes, V Shanks (1999) Confirmation of conjugation processes during TNT metabolism by axenic plant roots. Environ Sci Technol 33: 446–452. Buckman AH, CS Wong, EA Chow, SB Brown, KR Solomon, AT Fisk (2006) Biotransformation of polychlorinated biphenyls (PCBs) and bioformation of hydroxylated PCBs in fish. Aquat Toxicol 78: 176–185. Burken JG, JL Schnoor (1997) Uptake and metabolism of atrazine by poplar trees. Environ Sci Technol 31: 1399–1406. Campbell LM, DCG Muir, DM Whittle, S Backus, RJ Nostrom, AT Fisk (2003) Hydroxylated PCBs and other chlorinated phenolic compounds in lake trout (Salvelinus namaycush) blood plasma from the Great Lakes region. Environ Sci Technol 37: 1720–1725. Chen TP, SA Woodin, DE Lincoln, CR Lovell (1996) An unusual dehalogenating peroxidase from the marine terebellid polychaete Amphitrite ornata. J Biol Chem 271: 4609–4612. Collier TK, LC Thomas, DC Malins (1978) Influence of environmental temprature on disposition of dietary naphthalene in coho salmon (Oncorhynchus kisutch): isolation and identification of individual metabolites. Comp Biochem Physiol 61C: 23–28. Cruz SM, MN Scott, AK Merritt (1993) Metabolism of [14C]metolachlor in blueguill sunfish. J Agric Food Chem 41: 662–668. Dady JM, SP Bradbury, AD Hoffman, MM Voit, DL Olson (1991) Hepatic microsomal N-hydroxylation of aniline and 4-chloroaniline by rainbow trout (Oncorhyncus mykiss). Xenobiotica 21: 1605–1620. de Maagd P G-J, AD Vethaak (1998) Biotransformation of PAHs and their carcinogenic effects in fish. Handbook Environ Chem 3J: 265–309. de Wolf W, W Seinen, LM Hermens (1993) Biotransformation and toxicokinetics of trichloroanilines in fish in relation to their hydrophobicity. Arch Environ Contam Toxicol 25: 110–117. Donnarumma L, G de Angelis, F Gramenzi, L Vittozzi (1988) Xenobiotic metabolizing enzyme systems in test fish. III. Comparative studies of liver cytosolic glutathione S-transferases. Ecotoxicol Environ Saf 16: 180–186. Driscoll SK, AE McElroy (1996) Bioaccumulation and metabolism of benzo[a]pyrene in three species of polychaete worms. Environ Toxicol Chem 15: 1401–1410. Egaas E, U Varanasi (1982) Effects of polychlorinated biphenyls and environmental temperature on in vitro formation of benzo[a]pyrene metabolites by liver of trout (Salmo gairdneri). Biochem Pharmacol 31: 561–566. Feroz M, AA Podowski, MAQ Khan (1990) Oxidative dehydrochlorination of heptachlor by Daphnia magna. Pest Biochem Physiol 36: 101–105. Funari E, A Zoppinki, A Verdina, G de Angelis, L Vittozzi (1987) Xenobiotic metabolizing enzyme systems in test fish. I. Comparative studies of liver microsomal monooxygenases. Ecotoxicol Environ Saf 13: 24–31. Garrison AW, VA Nzengung, JK Avanta, JJ Ellington, WJ Jones, D Rennels, NL Wolfe (2000) Phytodegradation of p,p′-DDT and the enantiomers of o,p′-DDT. Environ Sci Technol 34: 1663–1670. Glickman AH, CN Statham, A Wu, JJ Lech (1977) Studies on the uptake, metabolism, and disposition of pentachlorophenol and pentachloroanisole in rainbow trout. Toxicol Appl Pharmacol 41: 649–658. Höhl H-U, W Barz (1995) Metabolism of the insecticide phoxim in plants and cell suspension cultures of soybean. J Agric Food Chem 43: 1052–1056. Huang KC, SF Collins (1962) Conjugation and excretion of aminobenzoic acid isomers in marine fishes. J Cell Comp Physiol 60: 49–52. Hunn JB, RA Schoettger, WA Willford (1968) Turnover and urinary excretion of free and acetylated MS 222 by rainbow trout, Salmo gairdneri. J Fish Res Bd Can 25: 215–231. James MO, JR Bend (1976) Taurine conjugation of 2,4-dichlorophenoxyacetic acid and phenylacetic acid in two marine species. Xenobiotica 6: 393–398.

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Janák K, G Becker, A Colmsjö, C Östman, M Athanasiadou, K Valters, Å Bergman (1998) Methyl sulfonyl polychlorinated biphenyls and 2,2-bis(4-chlorophenyl)-1,1-dichloroethene in gray seal tissues determinated by gas chromatography with electron capture detection and atomic emission detection. Environ Toxicol Chem 17: 1046–1055. Just CL, JL Schnoor (2004) Phytophotolysis of hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) in leaves of reed canary-grass. Environ Sci Technol 38: 290–295. Keizer J, G d′Agostino, R Nagel, F Gramenzi, L Vittozzi (1993) Comparative diazinon toxicity in guppy and zebra fish: different role of oxidative metabolism. Environ Toxicol Chem 12: 1243–1250. Kirchin MA, A Wiseman, DR Livingstone (1992) Seasonal and sex variation in the mixed-function oxygenase system of digestive gland microsomes of the common mussel, Mytilus edulis L. Comp Biochem Physiol 101C: 81–91. Koenig DW, HB Ward (1983) Prototheca zopfii Krüger strain UMK-13 growth on acetate or n-alkanes. Appl Environ Microbiol 45: 333–336. Krause A, WG Hancock, RD Minard, AJ Freyer, RC Honeycutt, HM LeBaron, DL Paulson, SY Liu, JM Bollag (1985) Microbial transformation of the herbicide metolachlor by a soil actinomycete. J Agric Food Chem 33: 584–589. LaCount MW, E Zhang, YP Chen, K Han, MM Whitton, DE Lincoln, SA Woodin, L Lebioda (2000) The crystal structure and amino acid sequence of dehaloperoxidase from Amphitrite ornata indicate common ancestry with globins. J Biol Chem 275: 18712–18716. Lambert M, S Kremer, H Anke (1995) Antimicrobial, phytotoxic, nematicidal, cytotoxic, and mutagenic activities of 1-hydroxypyrene, the initial metabolite in pyrene metabolism by the basidiomycete Crinipellis stipitaria. Bull Environ Contam Toxicol 55: 251–257. Larsen OFA, IS Kozin, AM Rija, GJ Stroomberg, JA de Knecht, NH Velthorst, C Gooijer (1998) Direct identification of pyrene metabolites in organs of the isopod Porcello scaber by fluorescence line narrowing spectroscopy. Anal Chem 70: 1182–1185. Letcher RJ, RJ Norstrom, DCG Muir (1998) Biotransformation versus bioaccumulation: sources of methyl sulfone PCB and 4,4′-DDE metabolites in polar bear food chain. Environ Sci Technol 32: 1656–1661. Lipke H, CW Kearns (1959a) DDT dehydrochlorinase I. Isolation, chemical properties, and spectrophotometric assay. J Biol Chem 234: 2123–2128. Lipke H, CW Kearns (1959b) DDT dehydrochlorinase II. Substrate and cofactor specificity. J Biol Chem 234: 2129–2132. Little PJ, MO James, JB Pritchard, JR Bend (1984) Benzo(a)pyrene metabolism in hepatic microsomes from feral and 3-methylcholanthrene-treated southern flounder, Paralichthys lethostigma. J Environ Pathol Toxicol Oncol 5: 309–320. Livingstone DR, SV Farrar (1984) Tissue and subcellular distribution of enzyme activities of mixed-function oxygenase and benzo[a]pyrene metabolism in the common mussel Mytilis edulis L. Sci Tot Environ 39: 209–235. Malins DC, BB McCain, DW Brown, MS Myers, MM Krahn, S-L Chan (1987) Toxic chemicals, including aromatic and chlorinated hydrocarbons and their derivatives, and liver lesions in white croaker (Genyonemus lineatus) from the vicinity of Los Angeles. Environ Sci Technol 21: 765–770. Malins DC, MM Krahn, MS Myers, LD Rhodes, DW Brown, CA Krone, BB McCain, S-L Chan (1985) Toxic chemicals in sediments and biota from a creosote-polluted harbor: relationships with hepatic neoplasms and other hepatic lesions in English sole (Parophrys vetulus). Carcinogenesis 6: 1463–1469. McElroy AE (1990) Polycyclic aromatic hydrocarbon metabolism in the polychaete. Nereis virens Aquat Toxicol 18: 35–50. McElroy AE, JC Means (1988) Uptake, metabolism, and depuration of PCBs by the polychaete. Nereis virens Aquat Toxicol 11: 416–417. Michel XR, PM Cassand, DG Ribera, J-F Narbonne (1992) Metabolism and mutagenic activation of benzo(a)pyrene by subcellular fractions from mussel (Mytilus galloprovincialis) digestive gland and sea bass (Discenthrarcus labrax) liver. Comp Biochem Physiol 103C: 43–51. Miyazawa M, T Wada, H Kameoka (1998) Biotransformation of (+) and (–) limonene by the larvae of common cutworm (Spodoptera litura). J Agric Food Chem 46: 300–303. Murphy SE, A Drotar, R Fall (1982) Biotransformation of the fungicide pentachloronitrobenzene by Tetrahymena thermophila. Chemosphere 11: 33–39. Neilson AH, H Blanck, L Förlin, L Landner, P Pärt, A Rosemarin, M Söderström (1989) Advanced hazard assessment of 4,5,6-trichloroguaiacol in the Swedish environment. In Chemicals in the Aquatic Environment (Ed L Landner), pp. 329–374. Springer, Berlin.

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Newman LA, SE Strand, N Choe, J Duffy, G Ekuan, M Ruiszai, BB Shurtleff, J Wilmoth, MP Gordon (1997) Uptake and biotransformation of trichloroethylene by hybrid poplars. Environ Sci Technol 31: 1062–1067. Niimi AJ, HB Lee, GP Kissoon (1989) Octanol/water partition coefficients and bioconcentration factors of chloronitrobenzenes in rainbow trout (Salmo gairdneri) Environ Toxicol Chem 8: 817-823. Nimmo IA (1987) The glutathione S-transferases of fish. Fish Physiol Biochem 3: 163–172. Olson LE, JL Allen, JW Hogan (1977) Biotransformation and elimination of the herbicide dinitramine in carp. J Agric Food Chem 25: 554–556. Pavlostathis SG, KK Comstock, ME Jacobson, FM Saunders (1998) Transformation of 2,4,6-trinitrotoluene by the aquatic plant Myriophyllum spicatum. Environ Toxicol Chem 17: 2266–2273. Peters R (1952) Lethal synthesis. Proc Roy Soc (London) B 139: 143–170 Pothuluri JV, JP Freeman, FE Evans, TB Moorman, CE Cerniglia (1993) Metabolism of alachlor by the fungus Cunninghamella elegans. J Agric Food Chem 41: 483–488. Roy S, O Hänninen (1994) Pentachlorophenol: uptake/elimination kinetics and metabolism in an aquatic plant Eichhornia crassipes. Environ Toxicol Chem 13: 763–773. Sandermann H (1994) Higher plant metabolism of xenobiotics: the “green liver” concept. Pharmacogenetics 4: 225–241. Schlenk D, DR Buhler (1991) Role of flavin-containing monooxygenase in the in vitro biotransformation of aldicarb in rainbow trout (Oncorhyncus mykiss). Xenobiotica 21: 1583–1589. Shang TQ, SL Doty, AM Wilson, WN Howald, MP Goprdon (2001) Trichloroethylene oxidative metabolism in plants: the trichloroethanol pathway. Phytochemistry 58: 1055–1065. Sijm DTHM, A Opperhuizen (1989) Biotransformation of organic chemicals by fish: enzyme activities and reactions. Handbook Environ Chem 2E: 164–235. Stapleton HM, B Brazil, RD Holbrook, CL Mitchelmore, R Benedict, A Konstatinov, D Potter (2006) In vivo and in vitro debromination of decabromodiphenmyl ether (BDE 209) by juvenile rainbow trout and common carp. Environ Sci Technol 40: 4653–4658. Stroomberg GJ, C Reuther, I Konin, TC van Brummelen, CAM van Gestel, C Gooijer, WP Cofino (1996) Formation of pyrene metabolites by the terrestrial isopod Porcello scaber. Chemosphere 33: 1905–1914. Ugrekhelidze D, F Korte, G Kvesitadze (1997) Uptake and transformation of benzene and toluene by plant leaves. Ecotoxicol Environ Saf 37: 24–29. van Aken B, JM Yoon, CL Just, JL Schnoor (2004) Metabolism and mineralization of hexahydro-1,3,5trinitro-1,3,5-triazine inside poplar tissues (Populus deltoides × nigra DN-34). Environ Sci Technol 38: 4572–4579. van Brummelen TC, NM van Straalen (1996) Uptake and elimination of benzo[a]pyrene in the terrestrial isopod Porcello scaber. Arch Environ Contam Toxicol 31: 277–285. Varanasi U, JE Stein, M Nishimoto, WL Reichert, TK Collier (1987) Chemical carcinogenesis in feral fish: uptake, activation, and detoxication of organic xenobiotics. Environ Health Perspect 71: 155–170. Walker JD, RS Pore (1978) Growth of Prototheca isolates on n-hexadecane and mixed-hydrocarbon substrate. Appl Environ Microbiol 35: 694–697. Walsh C (1982) Fluorinated substrate analogs: routes of metabolism and selective toxicity. Adv Enzymol 55: 187–288. Wei LY, CR Vossbrinck (1992) Degradation of alachlor in chironomid larvae (Diptera: Chironomidae). J Agric Food Chem 40: 1695–1699.

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Mechanistic Aspects

PART 1 INTRODUCTION OF OXYGEN This section addresses the cardinal issue of the mechanisms whereby oxygen is introduced into the substrates during degradation. A few comments are added on haloperoxidases, even though their major significance lies in the biogenesis of natural halogenated metabolites of bacteria, yeasts, and algae. Details of the reactions involved in degradation are covered in the appropriate parts of Chapters 7 through 11. There are several mechanisms by which oxygen is introduced during degradation: (a) molecular oxygen (dioxygen) may supply one atom of oxygen (monooxygenation), (b) dioxygen may supply both atoms of oxygen (dioxygenation), (c) the oxygen may originate from H2O (hydroxylation or hydration), and (d) by the activity of oxidases and peroxidases. A summary of oxygenation systems is given in Figure 3.1 and additional details of dioxygenases are given in Figure 3.2. The enzymes that introduce oxygen from water have been termed hydroxylases, but are more correctly termed oxidoreductases. These are particularly important in the degradation of azaarenes under both aerobic and anaerobic conditions.

MONOOXYGENATION Monooxygenases belong to a family of enzymes that introduce one atom of dioxygen into a wide range of substrates. These include aliphatic hydrocarbons and ketones, arene hydrocarbons, phenols, and hydroxybenzoates. For degradation of the aromatic compounds fission of the ring mediated by dioxygenases is necessary. These are discussed later in this section, and a comparison of the activities of monooxygenases and dioxygenases is given in Figure 3.3. Hydroxylation of Alkanes The oxidation of the simplest alkane methanol is carried out by methylotrophs that may be obligate or facultative. Only brief mention is made here of methane monooxygenase (MMO), which catalyzes the introduction of oxygen into methane. The enzyme exists in both a soluble (sMMO) and a particulate (pMMO) form of which the former has been more extensively studied. The enzyme consists of three components: a hydroxylase, a regulatory protein that is not directly involved in electron transfer between the hydroxylase, and a third protein that is a reductase containing flavin adenine dinucleotide (FAD) and a [2Fe–2S] cluster. Details of the structure of the hydroxylase and the mechanism of its action involving the FeIII–O–FeIII at the active site are given in a review (Lipscomb 1994). The particulate enzyme contains copper (Nguyen et al. 1994) or both copper and iron (Zahn and DiSpirito 1996), and the concentration of copper determines the catalytic activity of the enzyme (Sontoh and Semrau 1998). Alkane hydroxylase activity for higher alkanes is distributed among a number of Gram-negative bacteria, and sequence analysis of proteins shows that these differ widely even among pseudomonads (Smits et al. 2002). The alkane hydroxylase system for Pseudomonas putida (P. oleovorans) GPO1, which has been most extensively studied, is composed of three components: an integral membrane alkane hydroxylase (AlkB), a soluble rubredoxin (AlkG), and a soluble NADH-rubredoxin reductase (AlkT). The first of these is a representative of a large nonheme diiron cluster (Shanklin et al. 1997), while the others are involved in electron transport. The alkane hydroxylase of Acinetobacter sp. strain ADP1 is formally similar in requiring three components (Ratajczak et al. 1998). The alkane hydroxylase belongs to a family of nonheme iron oxygenases. There is some structural similarity between the nucleotide sequence of the integral membrane alkane hydroxylase and 103

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Environmental Degradation and Transformation of Organic Chemicals Source of oxygen O2  oxygenases 1.1

Monooxygenases

1.2

Dioxygenases

2

Source of oxygen H 2O  hydroxylases

2.1

Oxidoreductases

2.2

Hydroxylating system: chemical reaction with cation radicals produced by peroxidases

1

Oxygenases

1.1

Monooxygenases Flavoproteins introducing oxygen at site adjacent to existing hydroxy group 4-hydroxybenzoate hydroxylase Anthranilate hydroxylase in yeasts Salicylate hydroxylase Hydrocarbon hydroxylases introducing oxygen into nonoxygenated rings Toluene, ph enanthrene, pyrene monooxygenases Hydrocarbon epoxidases Fungal and yeast biotransformations of PAHs : production of phenols by NIH shift

1.2

Dioxygenases Dihydroxylating enzymes (Produce cis dihydrodiols: soluble multicomponent enzymes requiring NAD(P)H as cofactor) Benzene 1,2-dioxygenase Toluene 2, 3-dioxygenase Naphthalene 1,2-dioxygenase Biphenyl 2,3-dioxygenase Ring-cleavage enzymes (Carry out ring fission and require no cofactor) Catechol 2,3-dioxygenase (extradiol cleavage to carboxyaldehydes) Catechol 1,2-dioxygenase (intradiol cleavage to dicarboxylates) Hydroxylases

2.1

Oxidoreductases Mo multicomponent enzymes carrying out addition of H2O followed by dehydrogenation e.g., anaerobic hydratases coupled to dehydrogenases

2.2

Hydroxylation systems Secondary reaction with cation radicals formed by peroxidases

FIGURE 3.1 Bacterial oxygenation systems (excluding cytochrome P450 systems). (From Neilson, A.H. and Allard, A.-S. Microbial metabolism of PAHs and heteroarenes, The Handbook of Environmental Chemistry, Vol. 3J, pp. 1–80, Springer, 1998. With permission.)

the subunits of the monooxygenase encoded by xylA and xylM in the TOL plasmid that are involved in hydroxylation of the methyl groups in toluene and xylene in Pseudomonas putida PaW1 (Suzuki et al. 1991). Epoxidation of Alkenes The first step in the aerobic degradation of alkenes is epoxidation. Epoxidation is then followed by several alternatives. In one of them, the epoxides may undergo carboxylation: the enzyme

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105 Two −componentenzyme, e.g., benzoate, 4-sulfobenzoate

Class I

(1)

Reductase  flavin + [2Fe-2S] redox center

(2)

Oxygenase

Three −co mponent enzyme, e.g., benzene and toluene

Class II

Class III

(1)

Reductase  flavoprotein

(2)

Ferrodoxin  Rieske type

(3)

Oxygenase

Three−component type, e.g., naphthalene (1)

Reductase  flavin + [2Fe−2S] redox center

(2)

[2Fe−2S] ferrodoxin

(3)

Oxygenase

Futher division may be made according to the FMN or FAD flavin r equirement, and the nature of the [2Fe−2S] protein

FIGURE 3.2 Classification of bacterial dioxygenases. (From Neilson, A.H. and Allard, A.-S. Microbial metabolism of PAHs and heteroarenes, The Handbook of Environmental Chemistry, Vol. 3J, pp. 1–80, Springer, 1998. With permission.) Monooxygenase (a)

RH+ O 2 + H + NAD(P)H

(b)

+ O 2 + H + NAD(P)H

ROH +H 2 O +NAD(P) O + H 2 O+ NAD(P)

Dioxygenase (c)

H OH NAD(P) OH +

+ O 2 + H + NAD(P)H

H CO 2 H OH

CO 2 H

+ O2

(d) OH

FIGURE 3.3

Intradiol

Extradiol

CO 2 H OH CHO

Summary of (a,b) monooxygenase, (c,d) dioxygenase reactions.

complex has been purified from Xanthobacter sp. strain Py2, and includes an NADPH: disulfide oxidoreductase and three carboxylation components (Allen and Ensign 1997; Allen et al. 1999) (Figure 3.4). Alkene monooxygenase from this strain is closely related to the aromatic monooxygenases, and is able to hydroxylate benzene, toluene, and phenol (Zhou et al. 1999). The alkane hydroxylase from Pseudomonas oleovorans is able to carry out both hydroxylation and epoxidation (Ruettinger et al. 1977). Cycloalkanone Monooxygenases Flavoprotein 1,2-monooxygenases are used for the insertion of an atom of oxygen into the ring that is the first step in the degradation of both cyclopentanone and cyclohexanone before hydrolysis of the resulting lactones and dehydrogenation (Figure 3.5). There are two types of cycloalkanone

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Environmental Degradation and Transformation of Organic Chemicals NADP

FA D

NADP

NADP

FA DH

FA D

FAD

S

S

S

S

S CoM

S

S O

H

S

S

CH2

C

CoMS

OH +

CH3

CH2

CH3

HO2C−CH2−COCH3

FIGURE 3.4

C

CH3 COCH3

Mechanism for carboxylation of acetone involving FAD, NADP, and coenzyme M.

O

O

O

CO 2 H (CH 2)4 CH 2 OH

FIGURE 3.5

Monooxygenation of cyclohexanone.

monooxygenase: one that is FAD–NADPH-dependent, whereas the other is FMN–NADH-dependent. There has been renewed interest in these reactions in view of their biotechnological importance (Iwaki et al. 2002). Cycloalkanone monooxygenase for substrates with more than seven carbon atoms is, however, different (Kostichka et al. 2001). These monooxygenases are also involved in the degradation of cyclic terpenoids, and the enzyme from Rhodococcus erythropolis strain DCL14 is able to catalyze insertions of oxygen into the cyclohexanones produced from both (+)-(4R)- and (−)-(4S)limonene by Rhodococcus erythropolis (van der Werf et al. 1999). In addition, the monooxygenase from this strain is able to insert an oxygen atom into 2-ketoterpenoids at both the 1,2 and 2,3 positions, and of a 3-ketoterpenoid at the 3,4 position (terpenoid numbering) (van der Werf 2000). Monooxygenation of Aromatic Compounds Although the bacterial degradation of aromatic compounds is most frequently initiated by dioxygenation, monooxygenases may be involved additionally or alternatively for the introduction of oxygen into the ring. An alternative pathway for the degradation of alkyl arenes involves oxidation of methyl substituents to carboxylates (Figure 3.6). This is also used for the degradation of 4-nitrotoluene by Pseudomonas sp. strain TW3. The ntnWCMAB* genes that encode the enzymes, which convert the substrate into 4-nitrobenzoate, are similar to those in the upper pathway of the TOL plasmid (Harayama et al. 1989): an alcohol dehydrogenase, benzaldehyde dehydrogenase, a two-component monooxygenase, and part of a benzyl alcohol dehydrogenase (James and Williams 1998). In this strain, however, the enzymes are chromosomal, and the benzyl alcohol dehydrogenase ntnB* differs from the corresponding xylB protein. The oxidation by strains of Pseudomonas putida of the methyl group in arenes containing a hydroxyl group in the para position is, however, carried out by a different mechanism. The initial step is dehydrogenation to a quinone methide followed by hydration (hydroxylation) to the benzyl alcohol (Hopper 1976) (Figure 3.7). The reaction with 4-ethylphenol is partially stereospecific (McIntire et al. 1984), and the enzymes that catalyze the first two steps are flavocytochromes (McIntire et al. 1985). The role of formal hydroxylation in the degradation of azaarenes is discussed in the section on oxidoreductases (hydroxylases).

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107

CH 3

CO 2H

CO 2 H

CH 3

CO 2H

CO 2 H CO 2 H

OH OH

OH

OH

FIGURE 3.6 Degradation of toluene by side-chain oxidation.

FIGURE 3.7

CH3

CH2

CH2 OH

CHO

OH

O

OH

OH

Degradation of 4-methylphenol by hydroxylation.

Reactions of Aromatic Hydrocarbons Concurrent synthesis of both mono- and dioxygenases has been shown: 1. Several strains of Pseudomonas sp. that were induced for toluene dioxygenase activity catalyzed the enantiomeric monooxygenation of indane to indan-1-ol, and indene to inden-1-ol and cis-indan-1,2-diol (Wackett et al. 1988). Purified naphthalene dioxygenase from a strain of Pseudomonas sp. catalyzed the enantiomeric monooxygenation of indane to indan-1-ol and the dehydrogenation of indane to indene (Figure 3.8) (Gibson et al. 1995). 2. By cloning genes for benzene/toluene degradative enzymes in Pseudomonas (Burkholderia) sp. strain JS150, it has been found that this strain carries genes for a monooxygenase in addition to those for a dioxygenase (Johnson and Olsen 1995). Initial products from the metabolism of toluene after dehydrogenation are, therefore, 3-methyl catechol produced by 2,3-dioxygenation, 4-methylcatechol by 4-monooxygenation, and both 3- and 4-methylcatechols by 2-monooxygenation (Johnson and Olsen 1997). 3. Sphingomonas yanoikuyae (Beijerinckia sp.) strain B1 metabolizes biphenyl by initial dioxygenation followed by dehydrogenation to 2,3-dihydroxybiphenyl. Cells of a mutant (strain B8/36) lacking cis-biphenyl dihydrodiol dehydrogenase were induced with 1,3-dimethylbenzene, and could transform dihydronaphthalene by three reactions: (i) monooxygenation to (+)-(R)-2-hydroxy-1,2-dihydronaphthalene, (ii) dioxygenation to (+)-(1R,2S)-cis-naphthalene dihydrodiol, and (iii) dehydrogenation to naphthalene followed by dioxygenation to (+)-(1R,2S)-cis-naphthalene dihydrodiol (Eaton et al. 1996). 4. The degradation of pyrene by a Mycobacterium sp. involves both di- and monooxygenase activities (Heitkamp et al. 1988). 5. Arene oxides can be intermediates in the bacterial transformation of aromatic compounds and initiate rearrangements (NIH shifts) (Dalton et al. 1981; Cerniglia et al. 1984; Adriaens 1994). The formation of arene oxides may plausibly provide one mechanism for the formation of nitro-substituted products during degradation of aromatic compounds when nitrate is present in the medium. This is discussed in Chapter 2. 6. Toluene monooxygenases provide alternatives to dioxygenation, and the ortho-monooxygenase in Burkholderia (Pseudomonas) cepacia G4 has been shown to be carried

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Environmental Degradation and Transformation of Organic Chemicals OH

FIGURE 3.8 Transformation of indane to 3-hydroxyindane and indene.

on a plasmid TOM (Shields et al. 1995). In Pseudomonas mendocina KR1, the degradation of toluene is initiated by a three-component toluene-4-monooxygenase that converts toluene into 4-methylphenol. The formation of an intermediate arene oxide is consistent with the observation of an NIH shift in which 68% of the deuterium in 4-[2H]-toluene is retained in the 4-methyl phenol (Whited and Gibson 1991). In contrast, the enzyme that oxidizes toluene-4-sulfonate to the benzyl alcohol in the first step of its degradation is a monooxygenase that consists of only two components: an [Fe–S] flavoprotein that serves as a reductase and an oxygenase (Locher et al. 1991b). 7. Xanthobacter sp. strain Py2 was isolated by enrichment on propene that is metabolized by initial metabolism to the epoxide. The monooxygenase that is closely related to aromatic monooxygenases is able to hydroxylate benzene to phenol before degradation, and toluene to a mixture of 2-, 3-, and 4-methylphenols that are not further metabolized (Zhou et al. 1999). 8. Toluene/o-xylene monooxygenase in P. stutzeri strain OX1 carried out successive monooxygenation of o-xylene (Bertoni et al. 1998), and the toluene-4-monooxygenase of Pseudomonas mendocina KR1 and toluene-3-monooxygenase of Ralstonia pickettii PKO1 can hydroxylate benzene, toluene, and o-xylene (Tao et al. 2004; Vardar and Wood 2004). Reactions of Hydroxybenzoates and Related Compounds 1. Hydroxybenzoates can undergo hydroxylation with or without concomitant loss of CO2. For example, salicylate → catechol + CO2 (salicylate-1-hydroxylase) (Figure 3.9a) (WhiteStevens et al. 1972) and 4-hydroxybenzoate → 1,4-dihydroxybenzene + CO2 [(4-hydroxybenzoate 1-hydroxylase (decarboxylating)] in Candida parapsilosis (Figure 3.9b) (Eppink et al. 1997). 2. Hydroxylation without elimination of CO2 is illustrated in the following examples: a. Salicylate → 2,5-dihydroxybenzoate (salicylate-5-hydroxylase) (Zhou et al. 2002) (Figure 3.10a). b. Anthranilate → 5-hydroxyanthranilate and 2,5-dihydroxybenzoate (anthranilate-5hydroxylase) in Nocardia opaca (Figure 3.10b) (Cain 1968); Anthranilate → 2,3-dihydroxybenzoate in Trichosporon cutaneum (Figure 3.10c) (Powlowski et al. 1987). c. 3-Hydroxybenzoate → 2,5-dihydroxybenzoate (3-hydroxybenzoate-6-hydroxylase) (Gao et al. 2005). d. 4-Hydroxybenzoate → 3,4-dihydroxybenzoate (4-hydroxybenzoate-3-hydroxylase) (Entsch and Ballou 1989).

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109 OH

CO2 H

OH

OH (a)

CO2 H

OH

OH

OH

(b)

FIGURE 3.9 Monoxygenation of (a) salicylate and (b) 4-hydroxybenzoate. CO2 H

CO2 H

OH

OH (a) HO

CO2 H NH2 CO2 H NH2

(b)

HO CO2 H OH

(c) OH CO2 H

CO2 H OH

(d) NH2

H2 N

(e) N

OH

N

OH

OH

FIGURE 3.10 Monooxygenation of (a) salicylate, (b,c) 2-aminobenzoate, (d) 3-aminobenzoate, and (e) 2-hydroxyquinoline.

e. 3-Aminobenzoate → 5-aminosalicylate (3-aminobenzoate-6-hydroxylase) (Figure 3.10d) (Russ et al. 1994). f. Naphth-2-oate → 1-hydroxynaphth-2-oate (Morawski et al. 1997), or by a dioxygenation and elimination. g. Degradation of 4-hydroxybenzoate in Haloarcula sp. D1 proceeds via 2,5-dihydroxybenzoate involving an NIH shift (Fairley et al. 2002). h. 2-Hydroxyquinoline → 2,8-dihydroxyquinoline (Figure 3.10e) (Rosche et al. 1997).

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Environmental Degradation and Transformation of Organic Chemicals HO 2 C

OH

HO 2C

CO 2 H

O

O

CH2 CO 2 H

CO 2 H

CO 2 H

OH

O

HO

O

FIGURE 3.11

CO 2 H

CO 2 H

CH2 CO 2H

+ CH3CO 2H

CO 2 H

CO 2 H

The β-ketoadipate pathway.

CO 2 H HO 2 C

CHO

(a)

CO 2H

OH OH

OH

CO 2 H

CO 2H

(b) OH OH

CH 3− CO −CO 2 H + H − CO 2 H

CHO

CHO

CO 2 H

CO 2 H

OH

CH 3− CO −CO 2 H + CH 3 − CHO

OH

FIGURE 3.12 Biodegradation of 3,4-dihydroxybenzoate mediated by (a) 4,5-dioxygenase in Pseudomonas acidovorans and (b) 2,3-dioxygenase in Bacillus macerans.

Degradation of the products of hydroxylation then involves fission of the catechols (or 2,5dihydroxybenzoates) that are produced. All three fission pathways have been observed for 3,4-dihydroxybenzoate: a. Intradiol 3,4-dioxygenation in Pseudomonas putida and the β-ketoadipate pathway (Figure 3.11). b. Extradiol 4,5-dioxygenation in the Pseudomonas acidovorans group (Figure 3.12a). c. Extradiol 2,3-dioxygenation in species of Bacillus (Figure 3.12b). When 1,4-dihydroxy compounds are produced by monooxygenation, these are degraded by the gentisate pathway (Figure 3.13) mediated by gentisate dioxygenase (Wergath et al. 1998). Monooxygenation of Phenols Phenols Lacking Halogen or Nitro Substituents Monooxygenases generally initiate the degradation of phenols and their ethers: 1. The 4-methoxybenzoate monooxygenase from Pseudomonas putida shows low substrate specificity. Although it introduces only a single atom of oxygen into 3-hydroxy- and 4-hydroxybenzoate, it accomplishes the conversion of 4-vinylbenzoate into the corresponding side-chain diol (Wende et al. 1989). 2. The metabolism of 4-hydroxybenzoate involves conversion into 3,4-dihydroxybenzoate by a hydroxylase that has been purified and characterized from a strain of Pseudomonas

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111 H O 2C

CO2H

H O 2C

C O 2H

CO2H C O 2H C O −C O 2H

HO OH

CO −C O 2 H

OH

OH OH

C O 2H

C O 2H

FIGURE 3.13

The gentisate pathway.

fluorescens (Howell et al. 1972) and from P. aeruginosa strain PAO1 (Entsch and Ballou 1989). The enzyme is a flavoprotein containing FAD and requires NADPH for activity. 3. The enzymes in a strain of Pseudomonas putida (Arunachalam et al. 1992) and in Escherichia coli strain W (Prieto and Garcia 1994), which introduce an oxygen atom into 4-hydroxyphenylacetate to produce 3,4-dihydroxyphenylacetate, are flavoprotein monooxygenases that require a further protein component (coupling protein) for activity. In the absence of this protein, oxidation of NADH produces H2O2 (Arunachalam et al. 1992). This phenomenon has also been observed with other oxygenases including salicylate hydroxylase, 2,4-dichlorophenol hydroxylase, and 2,4,6-trichlorophenol monooxygenase. 4. Salicylate is an intermediate in the metabolism of PAHs including naphthalene and phenanthrene, and its degradation involves oxidation to catechol. The hydroxylase (monooxygenase) has been extensively studied (references in White-Stevens and Kamin 1972) and in the presence of an analog that does not serve as a substrate, NADH is oxidized with the production of H2O2 (White-Stevens and Kamin 1972). This “uncoupling” is characteristic of flavoenzymes and is exemplified also by the chlorophenol hydroxylase from an Azotobacter sp. that is noted later. 5. The degradation of 2-hydroxybiphenyl by Pseudomonas azelaica HBP1 is initiated by 2-hydroxybiphenyl 3-monooxygenase (Suske et al. 1999). Chlorophenols Monooxygenases carry out the introduction of oxygen, and in the degradation of 2,4,6-trichlorophenol by Ralstonia eutropha (Louie et al. 2002) and Burkholderia cepacia (Gisi and Xun 2003), they are FADH2-dependent (Gisi and Xun 2003). There are two situations that may pragmatically be distinguished: (a) ring fission of the chlorocatechols without direct loss of chlorine and (b) reactions that result in dechlorination before ring fission. Further details are given in Chapter 9, Part 2. 1. The hydroxylase that converts 2,4-dichlorophenol into 3,5-dichlorocatechol (Figure 3.14a) before ring fission has been purified from a strain of Acinetobacter sp. (Beadle and Smith 1982), and from Alcaligenes eutrophus JMP 134 (Don et al. 1985; Perkins et al. 1990). The reductant is NADPH, the enzyme is a flavoprotein containing FAD, and in the presence of compounds that are not substrates, NADPH and O2 are consumed with the production of H2O2.

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Environmental Degradation and Transformation of Organic Chemicals OH

OH

OH

Cl

CO 2 H CO 2H

(a) Cl

Cl

Cl

HO 2C

O

Cl

O Cl

Cl

HO 2C

HO 2 C

O

HO 2 C

O CO 2H

CO 2H

O CO 2H

Cl OH

OH

OH

OH

Cl

Cl

Cl

Cl

Cl

OH

Cl

Cl

Cl

Cl

Cl

Cl

OH

(b)

Cl

FIGURE 3.14

OH

OH

OH

Degradation of (a) 3,5-dichlorophenol and (b) pentachlorophenol.

2. A range of mechanisms operates for the degradation of chlorophenols with three or more substituents: a. The first step in the degradation of pentachlorophenol by Flavobacterium sp. 39732 involves introduction of oxygen at C-4 to produce tetrachlorohydroquinone (Figure 3.14b). The enzyme is not a hydroxylase, but a flavin-containing monooxygenase that can accept a number of halogenated phenols including 2,4,6-triiodophenol (Xun et al. 1992). b. An NADH-requiring chlorophenol monooxygenase from Burkholderia cepacia AC1100 successively dehalogenates 2,4,5-trichlorophenol to 2,5-dichloro-hydroquinone and 5-chloro-1,2,4-trihydroxybenzene (Xun 1996). It is a two-component enzyme, and component A contains FAD and an NADH reductase. c. The enzyme from Azotobacter sp. strain GP1, which catalyzes the formation of 2,6-dichlorohydroquinone from 2,4,6-trichlorophenol, is also a monooxygenase that requires NADH, FAD, and O2 (Wieser et al. 1997). The enzyme is able to accept other chlorophenols with consumption of NADH including 2,4-, 2,6-, 3,4-dichloro-, 2,4,5-trichloro-, and 2,3,4,5- and 2,3,4,6-tetrachlorophenol, and in the absence of a substrate results in unproductive formation of H2O2. d. The monooxygenase involved in the degradation of 2,4,6-trichlorophenol to 6-chloro2-hydroxyhydroquinone by Ralstonia eutropha strain JMP134 combines monooxygenation with hydrolytic activity, and it was shown that oxygen is introduced from both O2 and H2O (Xun and Webster 2004). Nitrophenols Monooxygenation is involved in the degradation of 4-nitrophenol by a strain of Moraxella sp. (Spain and Gibson 1991), and the enzyme appears to be a particulate flavoprotein (Figure 3.15a). In the Gram-positive organisms Arthrobacter sp. (Jain et al. 1994) and Bacillus sphaericus strain JS905 (Kadiyala and Spain 1998), 4-nitrophenol degradation involves initial hydroxylation to 4-nitrocatechol followed by elimination of nitrite with the formation of benzene-1,2,4-triol before ring fission

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113 OH

O

NO 2 O

(a)

(b)

OH

NO 2

OH OH

FIGURE 3.15

CO2H CO2H

OH

O

OH

O

OH

CO2H C HO

OH

OH O

OH

Alternative degradation of 4-nitrophenol.

CH 2 CH(NH 2 )CO 2 N

CH2CO−CO2H

CH2CO−CO2H OH

HO OH

FIGURE 3.16

Hydroxylation of phenylalanine followed by rearrangement to 2,5-dihydroxyphenylpyruvate.

to 3-ketoadipate (Figure 3.15b). A single two-component enzyme that consists of a flavoprotein reductase and an oxygenase carries out the first two reactions in B. sphaericus. Other Hydroxylations/Monooxygenations 1. Phenylalanine is hydroxylated to tyrosine and then sequentially to 4-hydroxyphenylpyruvate and by dioxygenation and rearrangement to 2,5-dihydroxyphenylpyruvate (Figure 3.16) (Arias-Barrau et al. 2004). Hydroxylation involves 6,7-dimethyltetrahydrobiopterin that is converted into the 4a-carbinolamine (Song et al. 1999). Copper is not a component of the active enzyme, although there is some disagreement on whether or not Fe is involved in the reaction for the hydroxylase from Chromobacterium violaceum (Chen and Frey 1998). 2. The enzyme in Rhodococcus sp. strain IGTS8 that brings about successive oxidation of dibenzothiophene to the sulfoxide and the sulfone is a flavin mononucleotide–dependent monooxygenase that carries out both reactions by sequential incorporation of a single atom of oxygen from O2 (Lei and Tu 1996). 3. The transformations of α- and β-endosulfan with the elimination of sulfite are carried out by flavin-dependent monooxygenation in Mycobacterium sp. (Sutherland et al. 2002) and Arthrobacter sp. (Weir et al. 2006).

CYTOCHROME P450 SYSTEMS These are found in bacteria, fungi, and yeasts, and mediate important hydroxylations. Reference should be made to reviews that discuss mechanistic aspects of cytochrome P450 reactions (Guengerich 1990), the reactions mediated by both eukaryotic and prokaryotic cytochrome

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Environmental Degradation and Transformation of Organic Chemicals Class I

Class II

Two-component enzyme of eukaryotic organisms (1)

NADPH reductase containing FAD and FMN

(2)

Monooxygenase

Three-component system of prokaryotic systems (1)

NADP reductase containing FAD

(2)

Ferrodoxin

(3)

Monooxygenase

FIGURE 3.17 Cytochrome P450 systems in prokaryotes and eukaryotes. (From Neilson, A.H. and Allard, A.-S. Microbial metabolism of PAHs and heteroarenes, The Handbook of Environmental Chemistry, Vol. 3J, pp. 1–80, Springer, 1998. With permission.)

P450 systems (Sariaslani 1991), their widespread role in the transformation of xenobiotics (Smith and Davis 1968), and their occurrence and activities in yeasts (Käppeli 1986). The essential features of prokaryotic and eukaryotic cytochrome P450 systems are compared in Figure 3.17. Heme oxygenase that carries out the degradation of heme after transport into the cell is important in a range of organisms, and some opportunistic pathogenic yeasts and bacteria have an enhanced requirement for Fe that can be satisfied by heme. In Gram-negative bacteria, heme is bound to outer membrane receptors and transported into the cell by the energy-transducing system TonB-ExcBExbD (Andrews et al. 2003; Wandersman and Delepelaire 2004). Heme oxygenases that have been designated HemO, HmuO, IsdG/I, and ChuZ mediate the degradation of heme with the release of Fe using an NADPH-cytochrome P450 reductase and three successive oxygenations: (a) hydroxylation at the α-meso position by the terminal oxygen of intermediate [Fe3+–O–OH], (b) release of CO and the production of verdoheme, and (c) further oxygenation to linear biliverdin with the release of Fe. Heme oxygenase has been identified in a number of pathogenic bacteria including Corynebacterium diphtheriae (Wilks and Schmitt 1998), Staphylococcus aureus (Skaar et al. 2004), species of Neisseria (Zhu et al. 2000), enterohemorragic Escherichia coli O157:H7 (Suits et al. 2005), and Campylobacter jejeuni (Ridley et al. 2006). In Pseudomonas aeruginosa, there are two heme oxygenases that differ at the sites of the heme that are initially hydroxylated to produce biliverdins IXβ and IXδ (Wegele et al. 2004). Although reaction with dioxygen of both cytochrome P450 and heme oxygenase involves the formation of a hydroperoxy Fe3+, the subsequent reactions are different, and activation of heme oxygenase is carried out by a nearby water molecule (Matsui et al. 2005). Further comments in the context of Fe utilization are given in Part 5 of this chapter. Prokaryotic Organisms The functioning of cytochrome P450 involves both oxygenation and reduction reactions (Tyson et al. 1972), and this is consistent with its role in both hydroxylation and reductive transformations. This activity is widely distributed among both Gram-negative and Gram-positive organisms, and mediates a number of important degradations and transformations including hydroxylations at quaternary carbon atoms. Cytochrome P450 systems in actinomyces have been reviewed (O′Keefe and Harder 1991), and the systematic nomenclature of some important bacterial cytochrome P450 systems has been given (Munro and Lindsay 1996). The diversity of reactions that are catalyzed by this system is illustrated by the following examples: 1. Hydroxylation of n-octane by cell extracts of Gordonia (Corynebacterium) sp. strain 7E1C (Cardini and Jurtshuk 1968), and of some strains of Acinetobacter calcoaceticus induced with n-hexadecane (Asperger et al. 1981). 2. Hydroxylation of long-chain carboxylic acids, amides, and alcohols—but not the esters or the corresponding alkanes—at the ω-1, ω-2, and ω-3 positions by a soluble enzyme system from Bacillus megaterium (Miura and Fulco 1975; Narhi and Fulco 1986, 1987).

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3. The transformation of benzene, toluene, naphthalene, biphenyl, and benzo[a]pyrene to the corresponding phenols (Trower et al. 1989) by Streptomyces griseus, and of phenanthrene by Streptomyces flavovirens to (−)trans-[9S,10S]-9,10-dihydrodihydroxyphenanthrene with minor amounts of 9-hydroxyphenanthrene (Sutherland et al. 1990). 4. The initial hydroxylation in the degradation of some terpenes: the ring methylene group of camphor by Pseudomonas putida (Katagiri et al. 1968; Tyson et al. 1972; Koga et al. 1986), and the isopropylidene methyl group of linalool by a strain of P. putida (Ullah et al. 1990). 5. When the 5 and 6 positions of camphor are blocked by substituents, hydroxylation at other positions may take place. For example, the quaternary methyl group of 5,5-difluorocamphor is hydroxylated to the 9-hydroxymethyl compound (Figure 3.18a) (Eble and Dawson 1984). 6. The cytochrome P450cam is able to bring about the stereoselective epoxidation of cismethylstyrene to the (1S,2R) epoxide (Ortiz de Montellano et al. 1991). 7. Adamantane (A) and adamantan-4-one (B) were specifically hydroxylated at the quaternary C-1 by cytochrome P450cam to produce C and D. In contrast, the eukaryotic cytochrome P450LM2 formed in addition to the C-2 compound from adamantane, and both 5-hydroxyadamantan-1-one (D) and the 4-anti-hydroxyadamantan-1-one (E) from adamantan-4-one (Figure 3.18b) (White et al. 1984). 8. The hydroxylation of cyclohexane by a strain of Xanthobacterium sp. (Trickett et al. 1990). In cell extracts, a range of other substrates was oxidized including cyclopentane, pinane, and toluene (Warburton et al. 1990). 9. Hydroxylation of progesterone and closely related compounds at the 15-β position by cell extracts of Bacillus megaterium (Berg et al. 1976). 10. The first step in the aerobic degradation of dehydroabietic acid by Pseudomonas abietaniphila strain BMKE-9 is hydroxylation at C-7 (Smith et al. 2004). 11. The reductive dehalogenation of polyhalogenated methanes (Castro et al. 1985) and polyhalogenated ethanes (Li and Wackett 1993) by Pseudomonas putida strain PpG786.

OH

(a)

O

O F

F

F

F

OH

(b)

A

B

O

C OH

OH

OH

D

FIGURE 3.18

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O

E

O

Hydroxylation of (a) 5,5-difluorocamphor, (b) adamantane (A) and adamantan-4-one (B).

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12. The oxidation of t-butyl methyl ether to t-butanol (Steffan et al. 1997), which is also mediated by the cytochrome P450 from camphor-grown Pseudomonas putida CAM, but not by that from Rhodococcus rhodochrous strain 116. 13. Initiation of the degradation of morpholine by a strain of Mycobacterium strain RP1 (Poupin et al. 1998), and additionally of pyrrolidine and piperidine. 14. The degradation of bisphenol-A by Sphingomonas sp. strain AO1 is initiated by hydroxylation to intermediates that undergo fission to 4-hydroxyacetophenone and 4-hydroxybenzoate. The components have been purified, and consist of cytochrome P450, ferredoxin reductase, and ferredoxin (Sasaki et al. 2005). 15. The transformation of sulfonylureas by Streptomyces griseolus grown in a complex medium containing glucose when the methyl group of the heterocyclic moieties is hydroxylated, and for some substrates, subsequently oxidized to the carboxylic acid (Romesser and O′Keefe 1986; O′Keefe et al. 1988). 16. The dealkylation of S-ethyl dipropylthiocarbamate and atrazine by a strain of Rhodococcus sp. (Nagy et al. 1995a,b), and of 2-ethoxyphenol and 4-methoxybenzoate by Rhodococcus rhodochrous (Karlson et al. 1993). In R. erythropolis NI86/21, however, the thcF gene involved in the degradation of thiocarbamate herbicides is a nonheme haloperoxidase that does not occur in other strains of rhodococci that can degrade thiocarbamates (de Schrijver et al. 1997): this specificity occurs also for the cytochrome P450 system encoded by the thcRBCD genes (Nagy et al. 1995a,b). Unusual aspects of bacterial cytochrome P450 systems are worth pointing out: a. The induction in B. megaterium by barbiturates, and the fact that the enzyme is unusual in having a single 119 kD protein-coupling NADH reduction to oxygenation (Narhi and Fulco 1986). b. The induction of the enzyme in S. griseus by genistein (5,7,4′-trihydroxyisoflavone) contained in the soybean flour that is used as the growth medium (Trower et al. 1989). c. Cytochrome P450 systems may be both substrate- and organism-specific, and multiple forms of the enzyme can be produced in the same organism. Eukaryotic Organisms Cytochrome P450 hydroxylation activity is well established in eukaryotic yeasts and some fungi, and this hydroxylation activity has attracted attention in various contexts. 1. The initial oxidation of aliphatic hydrocarbons and amines by yeasts (references in Käppeli 1986). The enzyme designated cytochrome P450aO is induced in organisms including species of Candida and Endomycopsis during growth with alkanes, and therefore differs from the enzyme (P45014DH) synthesized during the growth of saccharophilic yeasts under conditions of low oxygen concentration. 2. The aryl hydroxylase of Saccharomyces cerevisiae that transforms benzo[a]pyrene to the 3- and 9-hydroxy compounds, and the 7,8-dihydrodiol (King et al. 1984). 3. In fungi that serve as models for mammalian transformation of xenobiotics (Ferris et al. 1976; Smith and Rosazza 1983). 4. The hydroxylation of steroids at various positions is carried out by a range of fungi, and probably involves the cytochrome P450 system (Breskvar and Hudnik-Plevnik 1977). The removal of angular methyl groups is important in the transformation of steroids and related compounds. In these reactions, the methyl group is oxidized to the aldehyde before fission in which the carbonyl group oxygen is retained in formate (or acetate), and one oxygen atom from dioxygen

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is incorporated into the hydroxyl group. A group of formally analogous reactions using 18O2 has confirmed the mechanisms: 1. The oxidative removal of the 14α-methyl group in lanosterol by Saccharomyces cerevisiae with concomitant production of the ∆14–15 bond (Aoyama et al. 1984, 1987) and loss of formate (Figure 3.19a) is plausibly facilitated by formation of the conjugated 8-9,14-15-diene. The enzyme designated cytochrome P45014DM is induced under “semianaerobic” conditions during growth with glucose, and is different from both the P450aO enzyme and from P45022DS that brings about ∆22-dehydrogenation (Hata et al. 1983). The mechanism has been examined in a constructed strain containing the 14α-demethylase from Candida albicans and involves reaction of the carbonyl group with FeIII–O–OH (Shyadehi et al. 1996). 2. The conversion of androgens into estrogens involves removal of the C10 methyl group and aromatization of the A ring of ∆4,5-3-ketosterols. This is accomplished by oxidation to the C10 methyl group to the aldehyde, followed by loss of formate and aromatization of the ring to a 3-hydroxy-desmethylsterol (Figure 3.19b) (Stevenson et al. 1988). 3. Cytochrome P45017a carries out comparable reactions for removal of the C17 side chain of pregnenolone, and two reactions have been described both of which involved loss of acetate—17α-hydroxylation and formation of the 17-keto compound, and direct formation of the ∆16,17-ene (Figure 3.19c) (Akhtar et al. 1994). Most eukaryotic organisms such as yeasts and fungi contain peroxisomes, and the degradation of long-chain alkanes by yeasts is carried out in two separate organelles: hydroxylation to alkanols in microsomes (Käppeli 1986), and oxidation to the alkanoic acid CoA-esters that are further metabolized in peroxisomes. Peroxisomes are induced during growth with alkanes and

(a)

C

+ H−CO2H

D CHO CHO

(b)

A

B

+

O

O

COCH3

O

OH COCH3 (c)

C

H−CO2H

HO

+

CH3 CO2 H

D

+

CH3 CO2 H

FIGURE 3.19 Cytochrome P450 monooxygenation of (a) lanosterol, (b) aromatization of ∆4,5-3-ketosterols, and (c) pregnenolone.

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are multipurpose organelles. Some of the salient features include the following (Tanaka and Ueda 1993), which differ from the pathway for bacterial degradation of alkanoic acids: 1. In Candida sp., degradation of the CoA-alkanoic esters to the alkenoic acid esters is catalyzed by an acyl-CoA oxidase and results in the production of H2O2 that is converted into O2 by catalase activity. The enzyme from C. tropicalis contains FAD (Jiang and Thorpe 1983), and in C. lipolytica carries out a stereospecific antielimination of hydrogen (Kawaguchi et al. 1980). 2. A bifunctional enol-CoA hydratase and 3-hydroxyacetyl-CoA dehydrogenase are used in the degradation of CoA-alkenoic esters to the β-keto acid. This is then degraded to acetylCoA and the lower alkanoate ester by 3-ketoacetyl CoA thiolase and acetyl-CoA thiolase. 3. With the exception of the acetyl-CoA thiolase, all these enzymes are located exclusively in the peroxisomes, whereas the enzymes that are involved in lipid synthesis are found in the microsomes and the mitochondrion.

DIOXYGENATION Dioxygenases Involved in Dihydroxylation of Arenes The aerobic degradation of aromatic compounds involves the introduction of oxygen into the rings as a prelude to ring fission. The most common initial reaction is dioxygenation in which both atoms of oxygen are introduced to produce cis-dihydrodiols. Catechols are formed from these either by dehydrogenation or by elimination of substituents before further dioxygenation and fission of the ring. The biochemistry, enzymology, and genetics have been enunciated in great detail. The following overview attempts to bring together the reactions for many of the compounds that are discussed in the appropriate parts of Chapter 8. Group I The dioxygenases that catalyze the initial introduction of molecular oxygen into aromatic hydrocarbons have been extensively studied, and are multicomponent enzymes that generally carry out three functions: (a) reduction of NAD by a reductase, (b) electron transport by a ferredoxin and transfer of electrons onto the substrate simultaneously with (c) the introduction of oxygen by a terminal oxygenase. In contrast, there are a number of two-component dioxygenases that consist of only a reductase and an oxygenase. These contain a flavin cofactor and a [2Fe–2S] center, but lack the ferredoxin component that is invariably required for aromatic hydrocarbon dioxygenation. Examples of these two-component dioxygenases include a. 4-sulfobenzoate-3,4-dioxygenase (Locher et al. 1991a), 2-aminobenzoate-1,2-dioxygenase (Mampel et al. 1999); b. o-phthalate 4,5-dioxygenase (Batie et al. 1987); c. benzoate 1,2-dioxygenase (Yamaguchi and Fujisawa 1982); d. 4-chlorophenylacetate-3,4-dioxygenase from Pseudomonas sp. strain CBS32 (Schweizer et al. 1987); e. 3-chlorobenzoate 3,4-dioxygenase (Nakatsu et al. 1995); f. 2-halobenzoate 1,2-dioxygenase in Pseudomonas cepacia strain 2CBS (Fetzner et al. 1992). The corresponding dioxygenase from Pseudomonas aeruginosa 142 is, however, a threecomponent enzyme (Romanov and Hausinger 1994). The main classes of bacterial dioxygenases have already been outlined in Figure 3.1. Generally the cis-dihydrodiols are dehydrogenated to the corresponding catechol. In some, dehydrogenation is not required and elimination of CO2H, NH3, Hal−, NO2−, and SO32− or exceptionally OH produces catechols directly. These are assigned to group II.

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Group II In these reactions, dehydrogenation is not required and formation of the catechol is attained with concomitant elimination of CO2, NH3, NO2−, SO32−, or halide (Figure 3.20). Illustrative examples include the following: 1. a. Amine groups in the conversion of aniline into catechol (Bachofer and Lingens 1975; Fukumori and Saint 1997). b. Both amine and carboxyl groups in the conversion of anthranilic acid into catechol by anthranilate dioxygenase (Eby et al. 2001; Chang et al. 2003; Taniuchi et al. 1964). c. 2-Aminobenzenesulfonate to catechol-3-sulfonate and ammonia (Mampel et al. 1999). However, that there are also anthranilate monooxygenases that have already been noted. 2. a. Carboxyl groups in the conversion of benzoate into catechol by benzoate dioxygenase (Neidle et al. 1991). b. The degradation of 4-methyl-o-phthalate by P. fluorescens strain JT701 takes place by analogous oxidative decarboxylation with the formation of 4-methyl-2,3-dihydroxybenzoate (Ribbons et al. 1984), and of terephthalate (p-phthalate) to 3,4-dihydroxybenzoate and CO2. 3. Chloride during the conversion of 1,2,4,5-tetrachlorobenzene into 3,4,6-trichlorocatechol (Sander et al. 1991), and both chloride and carboxyl during the degradation of 2-chlorobenzoate (Fetzner et al. 1992; Romanov and Hausinger 1994). 4. Fluoride during the conversion of 3-fluorotoluene into 3-methylcatechol (Renganathan 1989), and of both fluoride and carboxyl during the degradation of 2-fluorobenzoate (Engesser et al. 1980). 5. Nitrite during conversion of 2-nitrotoluene into catechol (An et al. 1994), 2,4-dinitrotoluene into 4-nitrocatechol (Spanggord et al. 1991), and 1,3-dinitrobenzene to 4-nitrocatechol (Dickel and Knackmuss 1991). 6. Sulfite during the conversion of naphthalene-1-sulfonate into 1,2-dihydroxynaphthalene (Kuhm et al. 1991), and 4-carboxybenzenesulfonate into 3,4-dihydroxybenzoate (4-carboxycatechol) (Locher et al. 1991). Phenols may exceptionally be formed by spontaneous loss of H2O, and this is apparently favored by the presence of the electron-attracting substituents in 4-nitrotoluene, pyridine, quinoline, and indole. a. Degradation of o-xylene by Rhodococcus sp. DK17 is initiated by the formation of an unstable o-xylene cis-3,4-dihydrodiol. This may be dehydrogenated to 3,4-dimethylcatechol followed by extradiol ring fission, or undergo dehydration to produce 2,3- and 3,4-dimethylphenol (Figure 3.21a) (Kim et al. 2004). X

OH

X = Cl,

NH 2 ,

OH H

SO 3 H

OH OH

O HO

C

OH OH Y

FIGURE 3.20

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Y = H, Cl, NH 2 , F

Dioxygenation with concomitant loss of substituents.

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Environmental Degradation and Transformation of Organic Chemicals C H3 CH3 C H3

C H3

OH C H3

CH3 (a)

OH

OH OH

H

OH

OH

H

C H3 + CH3

CH3 CH3

CH3

CH3

CH3 OH

OH

H H

(b) N

N

C H3

N

OH

CH3

C H3

C H3 OH

OH

OH

H H

(c)

OH

OH NO2

N O2

NO2

NO 2

OH OH NH2

N OH H H OH

(d) N H

OH N H

H H

OH

N H

OH

FIGURE 3.21 (a) Degradation of 1,2-dimethylbenzene by dioxygenation and elimination with formation of dimethylphenols, (b) dioxygenation of 4-methylpyridine followed by elimination to 3-hydroxy-4-methylpyridine, (c) dioxygenation of 4-nitrotoluene followed by dehydrogenation or elimination, and (d) degradation of carbazole by dioxygenation followed by elimination.

b. 4-Methylpyridine → 3-hydroxy-4-methylpyridine by dioxygenation and spontaneous dehydration (Figure 3.21b) (Sakamoto et al. 2001). c. 4-Nitrotoluene → 2-methyl-5-nitrophenol + 3-methyl-6-nitrocatechol (Figure 3.21c) (Robertson et al. 1992). d. Quinoline → 3-hydroxyquinoline (Boyd et al. 1987). e. Quinol-2-one → 8-hydroxyquinol-2-one (Rosche et al. 1995). f. Degradation of carbazole by dioxygenation (Figure 3.21d) (Gieg et al. 1996).

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Arene hydrocarbon dioxygenases are capable of carrying out a number of reactions other than the introduction of both atoms of oxygen into the substrate. Illustrative examples of monooxygenation carried out by dioxygenases include the following: 1. Although naphthalene dioxygenase is enantiomer-specific in producing the (+)-cis-(1R,2S)dihydrodiol, it also possesses dehydrogenase and monooxygenase activities (Gibson et al. 1995), so that monooxygenase and dioxygenase activities are not exclusive. a. Naphthalene dioxygenase carried out dioxygenation of indene to cis-(1R,2S)-indandiol, monooxygenation to 1S-indenol, and in addition dehydrogenation of indane to indene (Lee et al. 1997). b. Toluene is oxidized by purified naphthalene dioxygenase from Pseudomonas sp. strain NCIB 9816-4 to benzyl alcohol and benzaldehyde, and ethylbenzene to (S)-1-phenylethanol and acetophenone. Whereas the initial reactions involve monooxygenation, oxidation to the aldehyde and ketone are dioxygen-dependent (Lee and Gibson 1996). c. Toluene dioxygenase from Pseudomonas putida strain F1 and chlorobenzene dioxygenase from Burkholderia sp. strain PS12 are involved in the degradation of toluene and chlorobenzene. In addition, they can carry out monooxygenation of 2- and 3-chlorotoluene by side chain oxidation to produce the corresponding benzyl alcohols that were slowly oxidized further (Lehning et al. 1997). 2. Naphthalene dioxygenase from P. putida strain F1 is able to oxidize a number of halogenated ethenes, propenes, and butenes, and cis-hept-2-ene and cis-oct-2-ene (Lange and Wackett 1997). Alkenes with halogen and methyl substituents at double bonds form allyl alcohols, whereas those with only alkyl or chloromethyl groups form diols. There are some metabolic similarities between naphthalene dioxygenase and 2,4-dinitrotoluene dioxygenase from Burkholderia sp. strain DNT, and it has been suggested that these enzymes may have a common ancestor (Suen et al. 1996). Protein sequences among biphenyl dioxygenases may be similar or identical, in spite of distinct differences in the range of PCB congeners that are attacked (Kimura et al. 1997; Mondello et al. 1997). The roles of the α- and β-subunits of the terminal dioxygenase (ISP) have been examined in more detail with divergent conclusions: a. Using enzymes from hybrid naphthalene dioxygenase and 2,4-dinitrotoluene dioxygenase genes introduced into Escherichia coli (Parales et al. 1998). Although the rates were different for the wild-type and hybrid enzymes, the products and the enantiomeric specificities were the same for each. It was shown that whereas the β-subunit of the dioxygenase was necessary for activity, it was the large α-subunit containing a Rieske [2Fe–2S] center that determined substrate specificity. b. Chimeras were constructed from the α- and β-subunits of the terminal dioxygenase proteins ISPBPH (Hurtubise et al. 1998) from two strains of PCB-degrading strains, Comamonas testosteroni strain B-356 and Pseudomonas sp. strain LB400. The enzymes were purified, and the substrate specificities of the purified enzymes showed that the structures of both subunits influenced the specificities. c. The function of the β-subunit that contains no detectable prosthetic groups is not fully understood, although the Tod C2 subunit is needed to obtain catalytic activity of the α-subunit (Tod C1) in Pseudomonas putida F1 (Jiang et al. 1999). It is clear from the preceding comments that there is no absolute distinction between the oxygenase activities mediated by dioxygenases. This is even less clear for heteroarenes than it is for carbocyclic compounds. An illustrative example is provided by Pseudomonas putida strain 86 in which 8-hydroxy-quinol-2-one is produced from quinol-2-one (Rosche et al. 1997).

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The enzyme system consists of a reductase that transfers electrons from NADH and contains FAD and a [2Fe–2S] ferredoxin, and a high-molecular-weight oxygenase consisting of six identical subunits and six Rieske [2Fe–2S] clusters. It was suggested that this complex belonged to class IB oxygenases that include benzoate-1,2-dioxygenase. It is therefore possible that this apparent monooxygenase is, in fact, a dioxygenase introducing oxygen at the 8,8a positions followed by elimination (Rosche et al. 1995). The formation of 1-hydroxynaphthalene-2-carboxylate from naphthalene-2-carboxylate was apparently not via (1R,2S)-cis-1,2-dihydro-1,2-dihydroxynaphthalene-2-carboxylate (Morawski et al. 1997). This could have taken place either by monooxygenation that is supported by the activity of naphthalene dioxygenase in which both activities are catalyzed by the same enzyme (Gibsonet al. 1995), or by an elimination that has already been noted. Exceptionally, an acyl chloride that is spontaneously hydrolyzed to the carboxylic acid may be formed: a. Pseudomonas putida strain GJ31 is able to grow with both toluene and chlorobenzene. Degradation of chlorobenzene proceeds by formation of the dihydrodiol and dehydrogenation to 3-chlorocatechol. This is degraded by a catechol 2,3-dioxygenase (extradiol or proximal) to produce the acyl chloride that is hydrolyzed to 2-hydroxymuconate (Figure 3.22a) (Mars et al. 1997; Kaschabek et al. 1998). b. The degradation of γ-hexachlorocyclohexane by Sphingomonas paucimobilis UT 6 takes place by several steps that result in the production of 2-chlorohydroquinone (Endo et al. 2005; Miyauchi et al. 2002). This is degraded by dioxygenation to an acyl chloride that is hydrolyzed to maleylacetate (Figure 3.22b) (Miyauchi et al. 1999). c. Sphingomonas chlorophenolica ATCC 39723 degrades pentachlorophenol in a series of steps to 2,6-dichlorohydroquinone that undergoes dioxygenation to an acyl chloride that is hydrolyzed to 2-chloromaleylacetate (Figure 3.22c) (Xun et al. 1999; Ohtsubo et al. 1999).

Cl

Cl OH

(a)

COCl CO2 H

CO2 H CO 2 H

CO2 H CO 2 H

OH

O

O

OH

Cl Cl

OH Cl

CO2 H COCl

Cl

(b) Cl

Cl Cl

OH

OH OH

OH Cl

Cl

Cl

Cl

Cl

Cl

Cl

(c)

Cl

CO2 H COCl

CO2 H CO 2 H

H

H OH

Cl

OH

O

FIGURE 3.22 Degradation involving dioxygenation via acyl chlorides of (a) chlorobenzene, (b) γ- hexachlorocyclohexane, and (c) pentachlorophenol.

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123

Ring-Fission Dioxygenases After dioxygenation and the formation of catechols, ring fission is mediated by a different group of dioxygenases. Group I Catechols and 2-Aminophenols

There are three possibilities, each of which has been realized (Figure 3.23), and the choice among them depends both on the substituents and on the organism. 1. Intradiol (or ortho) fission by catechol 1,2-dioxygenase breaks the bond between the atoms bearing the hydroxyl groups: a. Catechol 1,2-dioxygenase contains Fe(III), and lacks heme and sulfur–iron components. The protein consists of α- and β-subunits with masses of 30 and 32 in the combinations (α,β), (α,α), or (β,β), and the molecular mass of the enzyme is 63 kD. b. 3,4-Dihydroxybenzoate-3,4-dioxygenase is an Fe(III) protein and consists of an aggregate of α- and β-subunits with masses of 22.2 and 26.6 in an (α,β)12 structure with a molecular mass of 587. 2. Extradiol (or meta) fission by catechol 2,3-dioxygenase breaks the bond between one of the hydroxyl groups and the adjacent nonhydroxylated carbon atom. Extradiol 2,3-fission enzymes have been divided into three major classes (Table 3.1) (Spence et al. 1996; Peng et al. 1998). OH OH

Cl

(a)

(c)

(b) OH CO 2 H CHO

OH

Cl

FIGURE 3.23 and (c) distal.

CO 2 H CO 2 H Cl

CO 2 H COCl

Alternative dioxygenation pathways for 3-substituted catechols: (a) extradiol, (b) intradiol,

TABLE 3.1 Classes of Extradiol Dioxygenases Class I II

III

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Organism Rhodococcus globerulus P6 Burkholderia cepacia LB 400 Pseudomonas pseudoalcaligenes KF 707 Pseudomonas sp. KKS 102 Sphingomonas paucimobilis SYK 6 Pseudomonas sp. CA 10 Escherichia coli C Escherichia coli Alcaligenes eutrophus JMP 222 Sphingomonas paucimobilis SYK 6

Substrate for Dioxygenation 2,3-Dihydroxybiphenyl 2,3-Dihydroxybiphenyl 2,3-Dihydroxybiphenyl 2,3-Dihydroxybiphenyl 3,4-Dihydroxybenzoate 4,5-dioxygenase 2′-Amino-2,3-dihydroxybiphenyl 3,4-Dihydroxyphenylacetate 2,3-dioxygenase 2,3-Dihydroxyphenylpropionate Catechol 2,3-dioxygenase I 2,2′,3-Trihydroxy-3′-methoxybiphenyl-5,5′-dicarboxylate

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Environmental Degradation and Transformation of Organic Chemicals R= R

R

O O

R

O

O

O

CH 2 CH(NH 2 )CO 2 H COR

O

(a) N

N H OH

O

(b) OH O

O OH

O

N H

NHCHO

OH O

CO 2 H

O OH

CO 2 H

O

FIGURE 3.24

Dioxygenation of (a) catechol and (b) tryptophan.

3. Distal fission: Although substituted catechols generally undergo extradiol fission, this may produce toxic metabolites, and can be circumvented by regioselective 1,6-dioxygenation. This is rather uncommon but it has been observed in a number of degradations: a. 2-Aminophenol mediated by a 1,6-dioxygenase in Pseudomonas pseudoalcaligenes JS 45 (Lendenmann and Spain 1996) and in Comamonas CNB-1 (Wu et al. 2005). b. 3-Chlorocatechol by Sphingomonas xenophaga BN6 (Riegert et al. 2001) and 3-fluorobenzoate by strain FLB300 (that does not belong to the genus Pseudomonas) (Engesser et al. 1990; Schreiber et al. 1980). 4. Tryptophan dioxygenase that mediates the conversion of tryptophan into N-formylkynurenine, contains noncovalently bound heme, and is active in the Fe(II) form. Oxygenation is initiated at C3 with the subsequent formation of a dioxetane that decomposes into products (Figure 3.24a) (Leeds et al. 1993). The reaction may formally be considered analogous to catechol extradiol fission dioxygenation (Figure 3.24b). 5. There are a few extradiol-fission dioxygenases that are dependent on metals other than Fe. Magnesium

3,4-Dihydroxyphenylacetate extradiol dioxygenase from Klebsiella pneumoniae strain M5a1 contains magnesium and is not activated by Fe2+ (Gibello et al. 1994). Manganese

a. The 3,4-dihydroxyphenylacetate-2,3-dioxygenase from Arthrobacter globiformis (Boldt et al. 1995) and the 2,3-dihydroxybiphenyl-1,2-dioxygenase from Bacillus brevis strain JFG8 (Que et al. 1981) are Mn(II) enzymes, and are neither activated by Fe(II) nor rapidly inhibited by H2O2. b. The 2,3-dihydroxybiphenyldioxygenase from Bacillus sp. strain JF8 that mediated fission of the biphenyl ring is unusual in being Mn(II)-dependent, and differs in structure from those of Burkholderia sp. strain LB400 and Pseudomonas paucimobilis strain KF707 (Hatta et al. 2003). Copper

This is an unusual component of dioxygenases but is contained in the dioxygenase from Aspergillus japonicus that brings about fission of the C ring of quercitin with the formation of carbon monoxide (Steiner et al. 2002). However, although copper can be bound quite tightly to tryptophan 2,3-dioxygenase, the evidence suggests that it probably does not play a role in the catalytic activity of the enzyme (Ishimura et al. 1980).

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125

Group II Hydroxybenzoates

The second large group of ring-fission dioxygenases are used for the degradation of ortho-dihydroxybenzoates including, 2,3-, 3,4-, and 4,5-dihydroxybenzoate. As for simple catechols, fission may be either intradiol or extradiol, and the choice between them depends on the organism: a. The 3,4-dihydroxybenzoate dioxygenase in Pseudomonas putida mediates intradiol fission to a muconic acid by an Fe(III) enzyme (Bull and Ballou 1981). b. The 3,4-dihydroxybenzoate in Pseudomonas testosteroni mediates extradiol fission to a muconic semialdehyde by an Fe(II) enzyme (Arciero et al. 1983). Group III Other Dioxygenases

This comprises a heterogeneous group of enzymes that is used for the degradation of substrates including gentisate, salicylate, and 1-hydroxynaphthalene-2-carboxylate by pathways that do not involve catechols: 1. 2,5-Dihydroxybenzoate (gentisate) 1,2-dioxygenase brings about ring fission between the carboxyl and hydroxyl groups, and produces maleylpyruvate without loss of the carboxyl group (Figure 3.13) (Harpel and Lipscomb 1990). It belongs to a different class of ringcleaving dioxygenases from those that have already been described, since the protein sequences of the gene product from Sphingomonas sp. strain RW5 revealed little or no similarities to those of either intradiol or extradiol dioxygenases (Wergath et al. 1998; Hintner et al. 2001). In contrast, the degradation of the analogous salicylate proceeds by hydroxylation to catechol with loss of carboxyl, followed by ring fission of the catechol. 2. There are a few dioxygenases that carry out fission of aromatic rings and do not have vicinal hydroxyl groups: a. The dioxygenase from Nocardioides sp. strain KP7 brings about ring fission of 1-hydroxynaphthalene-2-carboxylate to 2′-carboxybenzalpyruvate, and is different from the dioxygenases that catalyze ring fission of catechols (Iwabuchi and Harayama 1998b). b. The dioxygenase from Pseudaminobacter salicylatoxidans is able to carry out direct ring fission of salicylates with a range of substituents including chlorine and bromine without the formation of intermediate catechols (Hintner et al. 2004). The deduced amino acid sequence encoded a protein to which the highest degree of similarity was found in a presumptive gentisate 1,2-dioxygenase. Whereas the ring fission products from 3-halosalicylates underwent decarboxylation to 6-halo-2-keto-3,5-hexa-dienoates, the corresponding products from 5-halosalicylates were lactones that were hydrolyzed to 4-hydroxy-2-keto-3,5-hexadiendioates. 3. The dioxygenase that brings about the degradation of 2-methyl-3-hydroxypyridine-5-carboxylate to α-(N-acetylaminomethylene)succinic acid (Sparrow et al. 1969). 4. The dioxygenase that catalyzes the degradation of 2,5-dihydroxypyridine to maleamate and formate (Gauthier and Rittenberg 1971). Group IV There is a rather large and important group of dioxygenases that do not bear any similarity to the dioxygenases already described: 1. One important group is dependent on Fe2+ and 2-ketoglutarate from which succinate and CO2 are produced concomitant with dioxygenation: a. The degradation of 2,4-dichlorophenoxyacetate by Alcaligenes eutrophus strain JMP134 involves initial formation of 2,4-dichlorophenol and glyoxylate. This is

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accomplished by a dioxygenase that couples this reaction to the conversion of 2-ketoglutarate into succinate and CO2 (Fukumori and Hausinger 1993a) (Figure 3.25a). The enzyme has been purified and characterized, and is able to accept other phenoxyacetates including 2,4,5-trichlorophenoxyacetate and 2-chloro-4-methylphenoxyacetates (Fukumori and Hausinger 1993b). In Burkholderia spp. that also degraded 2,4-dichlorophenoxyacetate by this pathway, the sequence of the chromosomal gene tfdA was, however, only 77.2% homologous to the plasmid-borne gene of A. eutrophus strain JMP 134. In contrast, it shared a 99.5% identity with the chromosomal gene from another strain of Burkholderia sp. from a geographically distinct area (Matheson et al. 1996). In Sphingomonas herbicidovorans, the 2-ketoglutaratedependent dioxygenase converted 4-chloro-2-methylphenoxypropionate (mecoprop) into 4-chloro-2-methylphenol, pyruvate, and succinate each of which were formed with incorporation of one atom of O2 (Nickel et al. 1997). b. A formally analogous reaction mediated by a dioxygenase is involved in the intramolecular dioxygenation of 4-hydroxyphenylpyruvate to 2,5-dihydroxyphenylacetate (Lindblad et al. 1970) in which both the 2-hydroxyl and the carboxylate oxygen atoms are derived from O2 (Figure 3.25b). The degradation of 4-hydroxyphenylpyruvate produced from tyrosine by Pseudomonas sp. strain PJ 874 grown at the expense of tyrosine is mediated by 4-hydroxyphenylpyruvate dioxygenase. This has been purified (Lindstedt et al. 1977; Johnson-Winters et al. 2003), and the primary structure of the pseudomonas enzyme has been determined (Rüetschi et al. 1992). The metabolism of l-tyrosine may proceed by transamination, dioxygenation to 4-phenylpyruvate followed by either ring fission, or oxidation to 2-carboxymethyl 1,4-benzoquinone and polymerization (Denoya et al. 1994). c. The degradation of 2-aminoethanesulfonate (taurine) by Escherichia coli takes place by α-hydroxylation at the position adjacent to the sulfonate group with the production of aminoacetaldehyde and sulfite, and is mediated by a 2-ketoglutarate-dependent dioxygenase (Eichhorn et al. 1997). The degradation of linear alkyl sulfate esters (C4 –C12) by Pseudomonas putida strain S-313 under sulfate limitation takes place by an analogous reaction to produce the corresponding aldehyde and sulfate. This enzyme is, however, able to use a wider range of 2-ketoacids (Kahnert and Kertesz 2000). 2. A rather heterogeneous group of reactions are carried out by dioxygenation: a. 3-Hydroxy-4-ketoquinoline dioxygenation → anthranilate + CO (Figure 3.26a) (Fischer et al. 1999). b. Quercitin dioxygenation → depside + CO (Figure 3.26b) (Krishnamurty and Simpson 1970) is mediated by a uniquely copper-dependent dioxygenase (quercitinase) (Steiner et al. 2002). c. In the methionine salvage pathway that has been elucidated in Klebsiella pneumoniae (Wray and Abeles 1995), a key intermediate is an enediol that undergoes dioxygenation. A single protein can carry out two oxygenations—the Fe-dependent enzyme carries out fission of the double bond, whereas the other enzyme that is Ni-dependent leads to the loss of CO and the formation of formate (Dai et al. 1999) (Figure 3.27). These reactions are formally analogous to the intradiol and extradiol fission of catechols. d. Dioxygenation of 2-chlorohydroquinone to produce 3-hydroxymuconate in Sphingomonas paucimobilis UT26 (Miyauchi et al. 1999, Endo et al. 2005). The product of dioxygenation is an acyl chloride that is hydrolyzed to a hydroxymuconate to avoid toxifying the dioxygenase. e. In the oxidation of n-propane to n-propanol by Nocardia paraffinicum (Rhodococcus rhodochrous) ATCC 21198, the ratio of hydrocarbon to oxygen consumed was 2:1 and suggests that the reaction of two molecules of propane and one of dioxygen produced

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Cl

CH 2

O

H

Cl

+

CO2 H

OH

O

CH−CO2H

C

CO2 H

CO2 H

O

O

O

O

CO 2

C

CH 2

+

HO

O

CO 2 H

CO2 H

O

O Cl

CO2 H

CH 2

O

Cl

OH

O O

HO

H O CO2 H

CH 2

OH

OH

CO2 H

O

Dioxygenation of (a) 2,4-dichlorophenoxyacetate and (b) 4-hydroxyphenylpyruvate.

O

O C

O

Cl

FIGURE 3.25

(b)

(a)

Cl

O

H

+

CHO−CO2H

OH

OH

CH2−CO2H

Mechanistic Aspects 127

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Environmental Degradation and Transformation of Organic Chemicals O OH

CO 2 H

(a)

+ CO N

NH 2 OH

OH

OH

OH HO

O

HO

O

O CO 2 H

(b) OH OH

OH

O

O R C

(c)

CHOH C

RCO 2 H + CH 3 CHO

OH

FIGURE 3.26

Dioxygenation of (a) 3-hydroxyquinol-4-one to anthranilate, (b) rutin, and (c) enediol. O R

O O H

O R

C

C

C

OH

O

OH

R

C

C

C

C

O CH(OH)

O R

C

OH

O

C

OH + H−CO2H

O

OH

O

O R

FIGURE 3.27

C

O H

O

O

C C O OH

O

CH(OH) R

C

OH

+

CO + H−CO2H

Dioxygenation of enediol to formate + CO in the methionine salvage pathway.

two molecules of alkanol (Babu and Brown 1984). This reaction is formally comparable to the oxidation of 2-nitropropane to acetone by flavoenzymes in Hansenula mrakii (Kido et al. 1978b) and Neurospora crassa (Gorlatova et al. 1998). f. The polyisoprenoid oxygenase (Braaz et al. 2004) in Xanthomonas sp. is extracellular and functions as a heme-dependent oxygenase with the production of 12-keto-4, 8-dimethyltrideca-4,8-diene-1-al (Braaz et al. 2004). g. An unusual reaction is involved in the oxygenation of a substituted 2-chloroaniline to the corresponding nitro compound (pyrrolonitrin) that is a metabolite of Pseudomonas fluorescens (Lee et al. 2005). The enzyme contains a Rieske [Fe–S] cluster and a mononuclear Fe center, and has a very limited range of anilines. 3. There are important dioxygenases from eukaryotic organisms: a. The oxidation of 2-nitropropane by the yeast Hansenula mrakii is carried out by a flavoenzyme and produces 2 mol of acetone from 2 mol of substrate and 1 molecule of O2 that is activated by conversion into superoxide (Kido et al. 1978b) (Figure 3.28). A similar enzyme for which 2-nitropropane is the optimal substrate has been purified and characterized from the heterothallic ascomycete Neurospora crassa (Gorlatova et al. 1998).

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Mechanistic Aspects

129 CH 3 CH 3

H C

CH 3 CH 3

C

O

NO 2

O

O

CH 3 CH 3

O

NO 2 H

C

CH 3 CH 3

H C NO2

2CH3COCH3

FIGURE 3.28 Dioxygenation of 2-nitropropane.

b. 1,2,4-Trihydroxybenzene is an intermediate in the degradation of a number of substrates by Phanerochaete chrysosporium including vanillate, 2,4-dichlorophenol, 2,4dinitrophenol, and 2,7-dichlorodibenzo[1,4]dioxin. Its degradation is mediated by a dioxygenase that carries out intradiol ring fission (Rieble et al. 1994). c. The degradation of rutin by Aspergillus flavus proceeds by hydrolysis to the aglycone followed by degradation to a depside with the release of unusual metabolite carbon monoxide (Figure 3.26b) is accomplished by dioxygenation (Krishnamurty and Simpson 1970). In addition, the dioxygenase is unusual in containing Cu in place of the more common Fe (Steiner et al. 2002). d. Several unusual 2-ketoglutarate-dependent dioxygenations have been found in the yeast Rhodotorula glutinis: i. Conversion of deoxyuridine by 1′-hydroxylation into uracil and ribonolactone (Stubbe 1985). ii. Thymine-7-hydroxylase catalyzed the oxidation of the 5-methyl group to 5-hydroxymethyl uracil (Wondrack et al. 1978).

INCORPORATION OF OXYGEN FROM WATER: OXIDOREDUCTASES AND HYDRATASES Oxidoreductases Both aerobic and anaerobic bacteria degrade azaarenes formally by hydroxylation in which the oxygen originates from water. The reaction involves a redox reaction and is mediated by oxidoreductases (dehydrogenases), although the physiological electron acceptors have not been generally established. These enzymes contain a molybdenum cofactor, and the formation of active xanthine dehydrogenase in purinolytic clostridia requires in addition a labile cofactor containing selenium. In contrast, the enzyme that mediates the hydroxylation of a number of bicyclic azaarenes has been termed an oxidase—quinaldine 4-oxidase—since oxygen functions effectively as an electron acceptor (Stephan et al. 1996). Five-Membered Heteroarenes The aerobic degradation of the 2-carboxylates of furan, pyrrole, and thiophene is initiated by hydroxylation before fission of the rings. Although details of the enzymes are limited, it was suggested on the basis of tungstate inhibition and 185W [tungstate] labeling that the degradation of 2-furoyl-coenzyme A involves a molybdenum-dependent dehydrogenase (Koenig and Andreesen 1990). Pyridine The aerobic degradation of pyridines with carboxyl and hydroxyl substituents generally involves hydroxylation (dehydrogenation). The dehydrogenases for the degradation of pyridine-2-carboxylate by Arthrobacter picolinophilus (Siegmund et al. 1990), and for two steps in the degradation of nicotine by Arthrobacter nicotinovorans (oxidans) (Freudenberg et al. 1988; Baitsch et al. 2001) contain molybdenum, flavin, and nonheme iron–sulfur. The nicotinate dehydrogenase from Clostridium barkeri contains, in addition, a labile selenium cofactor (Dilworth 1982).

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Quinoline The quinoline enzymes have been examined in much greater detail than those for pyridines, and occur in both Gram-negative and Gram-positive organisms. The oxidoreductases have been purified from a number of organisms that degrade quinoline. These include P. putida strain 86 (Bauder et al. 1990), Comamonas testosteroni strain 63 (Schach et al. 1995), Rhodococcus sp. strain B1 (Peschke and Lingens 1991), and Agrobacterium sp. strain 1B that degrades quinoline-4-carboxylate (Bauer and Lingens 1992). They have a molecular mass of 300–360 kDa and contain per molecule, eight atoms of Fe, eight atoms of acid-labile S, two atoms of Mo, and two molecules of FAD. The organic component of the pterin molybdenum cofactor is generally molybdopterin cytosine dinucleotide (Hetterich et al. 1991; Schach et al. 1995). The oxidoreductase that catalyzes the hydroxylation of 2-methylquinoline is able, in addition, to accept a range of other six-membered azaarenes (Stephan et al. 1996). Isoquinoline The degradation of isoquinoline by Pseudomonas diminuta strain 7 is initiated by an oxidoreductase that contains [2Fe–2S] centers and the cofactor molybdopterin cytosine dinucleotide (Lehmann et al. 1994). Purines Xanthine dehydrogenase that mediates the conversion of hypoxanthine into xanthine and uric acid has been studied extensively since it is readily available from cow’s milk. It has also been studied (Leimkühler et al. 2004) in the anaerobic phototroph Rhodobacter capsulatus, and the crystal structures of both enzymes have been solved. Xanthine dehydrogenase is a complex flavoprotein containing Mo, FAD, and [2Fe–2S] redox centers, and the reactions may be rationalized (Hille and Sprecher 1987): R−H + H2O + MoVI=S → ROH + MoIV–SH + H+ Some additional issues are worth noting: a. Xanthine dehydrogenase from the anaerobic Clostridium purinolyticum, Cl. acidi-urici, and Cl. cylindrosporum contains a labile selenium cofactor. b. In Clostridium purinolyticum, purine hydroxylase, which is a separate enzyme from xanthine dehydrogenase, hydroxylates purine to hypoxanthine (6-hydroxypurine) and xanthine (2,6-dihydroxypurine), which is then further hydroxylated to uric acid (2,6,8trihydroxypurine) by xanthine dehydrogenase (Self 2002). Like the xanthine dehydrogenases, it contains a selenium cofactor. c. Although the mammalian enzyme is synthesized in the dehydrogenase form XDH, it can readily be converted by oxidation of sulfuryl groups into the oxidase XO that exclusively uses O2 as its substrate. Hydratases These enzymes catalyze the addition of the elements of water to carbon–carbon double bonds (C=C), carbon–carbon triple bonds (C≡C), carbon–nitrogen double bonds (C=N), or carbon– nitrogen triple bonds (C≡N). These reactions are completely different from oxidoreductases since no redox reactions are involved. Illustrative examples include the following: C=C a. An alternative to monooxygenation is realized in the hydration of the substituted cyclohexanone derived from the fission product of 1,2-dihydroxynaphthalene during the degradation of 1,2,3,4-tetrahydronaphthalene (tetralin) (Figure 3.29) (Hernáez et al. 2002).

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131 OH

OH OH

HO

O

OH HO

O O CO 2 H

OH OH

OH

O

FIGURE 3.29 Aerobic degradation of tetralin.

(a)

H

C

C

(b)

R

N

C

H

H2 C

CHOH

CH 3 CHO

R−NH−CHO

FIGURE 3.30 (a) Hydration of acetylene to acetaldehyde and (b) transformation of an isonitrile to N-alkylformamide.

b. In addition to the oxidative pathway via (4-hydroxy-3-methoxybenzoyl)acetyl-ScoA, the degradation of feruloyl-SCoA to vanillin can be carried out by the alternative hydration followed by loss of acetate that has been found in Pseudomonas fluorescens (Gasson et al. 1998). c. The degradation of phenanthrene by Nocardioides sp. strain KP7 is initiated by dioxygenation to 1-hydroxynaphthalene-2-carboxylate, which is further dioxygenated to 2′-carboxybenzalpyruvate before hydration and aldolase fission to 2-carboxybenzaldehyde (Iwabuchi and Harayama 1998a). d. One step in the anaerobic degradation of benzoyl-CoA by the denitrifying Thauera aromatica involves hydration to cyclohexa-1,5-dienecarboxylate thioester to produce the 6-hydroxy compound (Boll et al. 2000). C≡C Acetylene hydratase from the anaerobe Pelobacter acetylenicus is a tungsten–iron–sulfur enzyme that resembles molybdopterin with W replacing Mo (Meckenstock et al. 1999), and catalyzes the addition of the elements of water to acetylene (Figure 3.30a). R–C≡N and R–N≡C a. Nitriles may be degraded to carboxylic acids either directly by the activity of a nitrilase in Bacillus sp. strain OxB-1 and Pseudomonas syringae B728a, or undergo hydration to amides followed by hydrolysis in P. chlororaphis (Oinuma et al. 2003). b. The hydratase in Pseudomonas putida N19-2 converts isonitriles into N-substituted formamides, and its activity is dependent on the presence of Cys-101 (Figure 3.30b) (Goda et al. 2002).

OXIDASES, PEROXIDASES, AND HALOPEROXIDASES Oxidases These are produced by both prokaryotes and eukaryotes, and catalyze a number of important reactions. They are flavoproteins that produce potentially destructive H2O2 that is removed by the activity of catalase or peroxidase. The reactions are formally outlined in Figures 3.31a through c.

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Environmental Degradation and Transformation of Organic Chemicals (a)

R2 CHX + H2 O + O2

(b)

RHC

R2 CO + HX + H2 O2

NH2 CO2 H (c)

FIGURE 3.31

RCO − CO2 H + NH3 + H2 O2

+ H2 O + O 2

RCH2 CH2 CO2 H + O2

RCH

CH−CO2 H + H2 O2

(a–c) Overview of reactions carried out by oxidases.

1. Primary amines are widely used as a nitrogen source by bacteria, and the fi rst step in their degradation involves formation of the corresponding aldehyde. Although this may be accomplished by dehydrogenation, some organisms use an amine oxidase. For example, whereas Pseudomonas putida (ATCC 12633) and P. aeruginosa (ATCC 17933) employ a dehydrogenase, Klebsiella oxytoca (ATCC 8724) and Escherichia coli (ATCC 9637) use a copper quinoprotein amine oxidase (Hacisalihoglu et al. 1997). The alternative flavoprotein oxidase is, however, used for deamination of tyramine by Sarcina lutea. The amine oxidase functions by oxidation of the amine to the aldehyde concomitant with the reduction of 2,4,5-trihydroxyphenylalanine quinone (TPQ) to an aminoquinol which, in the form of a Cu(I) radical reacts with O2 to form H2O2 Cu(II) and the imine. Methylamine dehydrogenase, however, involves tryptophylquinone (TTQ) but does not result in formation of H 2O2 (references in Stubbe and van der Donk 1998). 2. The aerobic degradation of nicotine produces an N-methylpyrrolidine as the first metabolite by dehydrogenation. This is then hydroxylated at the benzylic carbon atom by an FAD-containing oxidase (Dai et al. 1968), and the γ-N-methylaminobutyrate that is produced by fission of the N-methylpyrolidine ring is demethylated by an oxidase to 4-aminobutyrate (Chiribau et al. 2004). 3. Quinaldine 4-oxidase from Arthrobacter sp. Rü61a that mediates the hydroxylation of a number of bicyclic azaarenes has been termed an oxidase since oxygen functions effectively as an electron acceptor (Stephan et al. 1996). 4. Hyphomicrobium sp. strain EG is able to grow at the expense of dimethyl sulfide or dimethyl sulfoxide (DMSO) and produces methanethiol initially. This is then further oxidized to formaldehyde, sulfide, and H2O2 by an oxidase that has been purified (Suylen et al. 1987). 5. The oxidation of cholesterol to cholest-4-ene-3-one is carried out by an oxidase in several bacteria. This activity has been found in Brevibacterium sterolicum and Streptomyces sp. strain SA-COO (Ohta et al. 1991), and the extracellular enzyme that has been purified from Pseudomonas sp. strain ST-200 (Doukyu and Aono 1998) has a preference for 3βhydroxy compounds. 6. d-Aminoacid oxidase has been isolated from a number of yeasts, and the nucleotide sequence of the enzyme from Rhodotorula gracilis ATCC 26217 has been established (Alonso et al. 1998). The gene could be overexpressed in Escherichia coli, and levels of the enzyme were greater under conditions of low aeration: the enzyme isolated from the recombinant organisms was apparently the apoenzyme since maximum activity required the presence of FAD. 7. Yeasts belonging to the genera Candida and Endomycopsis are able to degrade alkanes. The initial hydroxylation is carried out by cytochrome P450 that is found in the microsome (references in Käppeli 1986). The degradation of the alkanoic acids is carried out by enzymes that are contained in peroxisomes, and are induced during growth with alkanes. Acyl-CoA oxidase carries out the first step in the degradation of the alkanoic acid CoA-esters, and although this is formally a dehydrogenase, acyl-CoA dehydrogenase

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activity is absent (Kawamoto et al. 1978; Tanaka and Ueda 1993). The three subsequent enzymatic activities are apparently contained in a single protein. 8. Nitroalkanes are degraded by some yeasts and fungi, and the dioxygenase pathway that has already been described is not the only one. The fungus Fusarium oxysporum is able to degrade a number of nitroalkanes including 1- and 2-nitropropane, nitroethane, and nitrocyclohexane by an inducible oxidase, and the reaction rate of the enzyme is enhanced by addition of catalase and inhibited by both NADH and NADPH (Kido et al. 1978a). The oxidase is a flavoprotein that occurs in an inactive form as 5-nitrobutyl-FAD that is convertible into the active form of the enzyme (Gadda et al. 1997). This enzyme is clearly different from the dioxygenase. 9. Vanillyl-alcohol oxidase that brings about oxidation to vanillin is mediated by a flavoprotein that is reduced with production of a quinonemethine that reacts with H2O to form the aldehyde. The reduced flavoprotein is then reoxidized to produce H2O2 (Fraaije and van Berkel 1997). The enzyme is also able to accept a range of 4-alkylphenols. Two reactions occurred, and their ratio depended on the alkyl substituent: for example, whereas 4-hydroxyphenylethane produced 4-hydroxyacetophenone (76%) and 4-hydroxyphenylethene (24%), only the corresponding alkenes were produced from substrates with C4 –C7 alkyl groups (van den Heuvel et al. 1998). The reaction is formally analogous to the bacterial reaction catalyzed by 4-ethylphenol methylene hydroxylase, which proceeds by dehydrogenation followed by hydroxylation with H2O (Hopper and Cottrell 2003). Peroxidases Extracellular H2O2 is required for the activity of peroxidases in white-rot fungi, and this can be produced by several fungal reactions: a. Glyoxal oxidase is produced from Phanerochaete chrysosporium under high concentrations of oxygen, is stimulated by Cu2+, and oxidizes a range of substrates with the production of H2O2 from O2. Substrates include methylglyoxal, glyoxylic acid, and glycolaldehyde, and the pure enzyme requires activation by lignin peroxidase: under these conditions, in the presence of catalytic amounts of H 2O2, pyruvate and veratraldehyde are produced, respectively, from methylglyoxal and 3,4-dimethoxybenzyl alcohol (veratryl alcohol) (the lignin peroxidase substrate) (Kersten 1990). b. An aryl-alcohol oxidase that is produced from Bjerkandera adusta optimally under carbon limitation oxidized a number of benzyl alcohols including 4-methoxybenzyl alcohol, 3,4-dimethoxybenzyl alcohol (veratryl alcohol), and 4-hydroxy-3-methoxybenzyl alcohol, with the production of H2O2 from O2: monosaccharides were not oxidized (Muheim et al. 1990). An aryl-alcohol oxidase from Pleurotus eryngii is a flavoprotein with range of substrates comparable to that from B. adusta (Guillén et al. 1992). c. Peroxide may also be produced by the manganese peroxide–dependent oxidation of glycolate and oxalate that are synthesized by Ceriporiopsis subvermispora (Urzúa et al. 1998). Two distinct extracellular enzymes are produced by white-rot fungi—lignin peroxidase (LiP) and manganese peroxidase (MnP). These are produced under specific growth conditions that include carbon, nitrogen or sulfur limitation, manganese concentration, and increased oxygen concentration. They are involved in the degradation of lignin and in the biotransformation of xenobiotics. These enzymes require extracellular H2O2, which is produced by a number of these organisms as a result of oxidase activity that has already been noted. Details of the relevant reactions have emerged from a study of the degradation of a model lignin substrate by Phanerochaete chrysosporium (Hammel et al. 1994). 1-(3,4-Dimethyoxyphenyl)-2-phenoxypropane-1,3-diol was metabolized in the presence of H2O2 to glycolaldehyde that was identified by 13C-NMR: the oxidase activity of glyoxal

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oxidase that was synthesized simultaneously with lignin peroxidase produced oxalate and 3 mol of H2O2 that could then be recycled. In the degradation of lignin and model compounds, lignin peroxidase functions by generating cation radicals from aromatic rings. This results in fission of the alkyl side chain between C1 and C2 (Kirk and Farrell 1987, or—for substrates such as PAHs—may be followed by nonenzymatic nucleophilic reactions of the cation radical (Hammel et al. 1986; Haemmerli et al. 1986). Manganese peroxidase oxidizes Mn(II) to Mn(III), which is the active oxidant, and in the presence of H2O2 is capable of oxidizing a number of PAHs (Bogan and Lamar 1996) and mineralizing substituted aromatic compounds (Hofrichter et al. 1998). Both activities may bring about oxidation of PAHs in reactions that are mimicked by oxidation with manganese(III) acetate (Cremonesi et al. 1989) or electrochemical oxidation (Jeftic and Adams 1970). Haloperoxidases These are of primary significance in the biosynthesis of organohalogen compounds (Neilson 2003), which are distributed among mammals, marine biota, bacteria, and fungi. These enzymes catalyze the reaction: S–H + H2O2 + HalH → S–Hal + 2H2O where Hal may be chloride, bromide, or iodide (references in Neidleman and Geigert 1986). Haloperoxidases have been isolated and purified from a number of organisms in which they mediate biosynthetic reactions. There are several structural groups of haloperoxidases that are summarized in Table 3.2. a. The chloroperoxidase from Caldaromyces fumago has been isolated in pure form. It is a glycoprotein containing ferroprotoporphyrin IX, and displays additionally both peroxidase and catalase activities. In the absence of organic substrates it catalyzes the formation of Cl2 and Br2 from chloride and bromide, respectively (Morris and Haeger 1966; Libby et al. 1982). b. Bromoperoxidase has been isolated from Pseudomonas aureofaciens ATCC 15926, also displays peroxidase and catalase activities, and contains ferriprotoporphyrin IX (van Pée and Lingens 1985b). Four different bromoperoxidases have been isolated from Streptomyces griseus (Zeiner et al. 1988). Only one of them, however, contains ferriprotoporphyrin IX, and displays peroxidase and catalase activities. This illustrates that there are different groups of enzymes, one of which lacks heme prosthetic groups. Chloroperoxidase activity has also been found among degradative enzymes: 1. An inducible enzyme in the bacterium Rhodococcus erythropolis NI86/21 that is involved in the degradation of thiocarbamate herbicides is a nonheme haloperoxidase, which does not occur in other strains of rhodococci that can degrade thiocarbamates (de Schrijver et al. 1997). 2. Acinetobacter calcoaceticus strain F46 was enriched with fluorene that is degraded by dioxygenation, loss of C3, decarboxylation, and Baeyer–Villiger oxygenation to 3,4-dihydrocoumarin (Figure 3.32). The hydrolase for this was able to brominate monochlorodimedone in the presence of H2O2 and 3,4-dihydrocoumarin, or acetate, or butyrate (Kataoka et al. 2000). It was proposed that acyl peroxides could be produced from an acylated serine site on the enzyme by the action of H2O2 and oxidize bromide to the active brominating agent BrO−. This is analogous to the mechanism proposed earlier (Picard et al. 1997).

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TABLE 3.2 Types of Haloperoxidases Type of Enzyme

Group

Organism

Halogen

Bacterium Streptomyces phaeochromogenes Rhodophyta Cystoclonium purpureum Chlorophyta Penicillus capitatus

Heme

Heme + f Flavin Vanadium

Nonheme/flavin Nonheme/ nonflavin Nonheme/ nonmetal

Fungus Polychaete Phaeophyta Phaeophyta Phaeophyta Phaeophyta Phaeophyta Rhodophyta Fungus Bacterium

Caldariomyces fumago Notomastus lobatus Ascophyllum nodosum Ecklonia stolonifera Laminaria digitata Laminaria saccharina Macrocystis pyrifera Corallina pilulifera Curvularia inequalis Pseudomonas fluorescens

Cl Cl Br Br Br Br Br Br Cl Cl

van Pée and Lingens (1985a) Pedersén (1976) Baden and Corbett (1980) and Manthey and Hager (1981) Morris and Hager (1966) Chen et al. (1991) Vilter (1984) Hara and Sakurai (1998) Jordan and Vilter (1991) de Boer et al. (1986) Soedjak and Butler (1990) Krenn et al. (1989) and Itoh et al. (1988) Simons et al. (1995) Keller et al. (2000)

Bacterium

Pseudomonas putida

Br

Itoh et al. (1994)

Bacterium

Pseudomonas pyrrocinia, Streptomyces aureofaciens, and Acinetobacter calcoaceticus

Cl, Br, and Br

Wiesner et al. (1988), Weng et al. (1991), and Kataoka et al. (2000)

OH

Br Br Br

Reference

OH

O

OH CHO

O

FIGURE 3.32

O

Degradation of fluorene.

3. The gene encoding the esterase from Pseudomonas fluorescens was expressed in Escherichia coli, and the enzyme displayed both hydrolytic and bromoperoxidase activity (Pelletier and Altenbuchner 1995).

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Weng M, O Pfeifer, S Kraus, F Lingens, K-H van Pée (1991) Purification, characterization and comparison of two non-heme bromoperoxidases from Streptomyces aureofaciens. J Gen Microbiol 137: 2539–2546. Wergath J, H-A Arfmann, DH Pieper, KN Timmis, R-M Wittich (1998) Biochemical and genetic analysis of a gentisate 1,2-dioxygenase from Sphingomonas sp. strain RW 5. J Bacteriol 180: 4171–4176. White RE, MB McCarthy, KD Egeberg, SG Sligar (1984) Regioselectivity in the cytochromes P-450: control by protein constraints and by chemical reactivities. Arch Biochem Biophys 228: 493–502. Whited GM, DT Gibson (1991) Separation and partial characterization of the enzymes of the toluene-4-monooxygenase catabolic pathway in Pseudomonas mendocina KR1. J Bacteriol 173: 3017–3020. White-Stevens RH, H Kamin, QH Gibson (1972) Studies of a flavoprotein, salicylate hydroxylase II Enzyme mechanism. J Biol Chem 247: 2371–2381. White-Stevens RH, H Kamin (1972) Studies of a flavoprotein, salicylate hydroxylase. I. Preparation, properties, and the uncoupling of oxygen reduction from hydroxylation. J Biol Chem 247: 2358–2370. Wieser M, B Wagner, J Eberspächer, F Lingens (1997) Purification and characterization of 2,4,6trichlorophenol-4-monooxygenase, a dehalogenating enzyme from Azotobacter sp. strain GP1. J Bacteriol 179: 202–208. Wiesner W, K-H van Pée, F Lingens (1988) Purification and characterization of a novel non-heme chloroperoxidase from Pseudomonas pyrrocinia. J Biol Chem 263: 13725–13732. Wilks A, MP Schmitt (1998) Expression and characterization of a heme oxygenase (Hmu O) from Corynebacterium diphtheriae. J Biol Chem 273: 837–841. Wondrack LM, C-A Hsu, MT Abbott (1978) Thymine-7-hydroxylase and pyrimidine deoxyribonucleoside 2′-hydroxylase activities in Rhodotorula glutinis. J Biol Chem 253: 6511–6515. Wray JW, RH Abeles (1995) The methionine salvage pathway in Klebsiella pneumoniae and rat liver. J Biol Chem 270: 3147–3153. Wu J-F, C-W Sun, C-Y Jiang, Z-P Liu, S-J Liu (2005) A novel 2-aminophenol 1,6-dioxygenase involved in the degradation of p-chloronitrobenzene by Comamonas strain CNB-1: purification, properties, genetic cloning and expression in Escherichia coli. Arch Microbiol 183: 1–8. Xun L (1996) Purification and characterization of chlorophenol 4-monooxygenase from Burkholderia cepacia AC1100. J Bacteriol 178: 2645–2649. Xun L, CM Webster (2004) A monooxygenase catalyzes sequential dechlorinations of 2,4,6-trichlorophnel by oxidative and hydrolytic reactions. J Biol Chem 279: 6696–6700. Xun L, E Topp, CS Orser (1992) Diverse substrate range of a Flavobacterium pentachlorophenol hydroxylase and reaction stoichiometries. J Bacteriol 174: 2898–2902. Xun L, J Bohuslavek, M Cai (1999) Characterization of 2,6-dichloro-p-hydroquoinone 1,2-dioxygenase (PcpA) of Sphingomonas chlorophenolica ATCC 39723. Biochem Biophys Res Commun 266: 322–325. Yamaguchi M, H Fujisawa (1982) Subunit structure of oxygenase components in benzoate 1,2-dioxygenase system from Pseudomonas arvilla C-1. J Biol Chem 257: 12497–12502. Zahn JA, AA DiSpirito (1996) Membrane associated methane monooxygenase from Methylococcus capsulatus (Bath). J Bacteriol 178: 1018–1029. Zeiner R, K-H van Pée, F Lingens (1988) Purification and partial characterization of multiple bromoperoxidases from Streptomyces griseus. J Gen Microbiol 134: 3141–3149. Zhou N-Y, A Jenkins, CKN Chan, KW Chion, DJ Leak (1999) The alkene monooxygenase from Xanthobacter strain Py2 is closely related to aromatic monooxygenases and catalyzes aromatic monooxygenation of benzene, toluene, and phenol. Appl Environ Microbiol 65: 1589–1595. Zhou N-Y, J Al-Dulayymi, MS Baird, PA Williams (2002) Salicylate 5-hydroxylase from Ralstonia sp. strain U2: a monooxygenase with close relationships to and shared electron transport proteins with naphthalene dioxygenase. J Bacteriol 184: 1547–1555. Zhu W, A Wilks, I Stojiljkovic (2000) Degradation of heme in Gram-negative bacteria: the product of the hemO gene of Neisseriae is a heme oxygenase. J Bacteriol 182: 6783–6790.

PART 2 ELECTRON ACCEPTORS OTHER THAN OXYGEN INTRODUCTION Bacteria can use a range of electron acceptors in the absence of oxygen. Although they are able to reduce a number of oxyanions, only a limited number of these can support growth under anaerobic conditions by coupling reduction to the production of energy by proton translocation. A number

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of facultatively anaerobic bacteria are, however, able to carry out a respiratory metabolism in the absence of oxygen using alternative inorganic electron acceptors. For example, the following reductions may be coupled to oxidative degradation: nitrate to nitrogen (or nitrous oxide), sulfate to sulfide, carbonate to methane, fumarate to succinate, trimethylamine-N-oxide (TMAO) to trimethylamine, or dimethylsulfoxide to dimethyl sulfide. Some introductory comments are given, and an attempt is made to illustrate the metabolic potential of organisms under these conditions. It is appropriate to note the metabolic potential of members of the DMSO reductase family (McEwan et al. 2002) that are able to carry out a wide range of dissimilations: Me2SO → Me2S (DMSO reductase); Me3NO → Me3N (TMAO reductase); HCO2H → CO2 + H2 (formate dehydrogenase); AsO2− → AsO43− (arsenite oxidase); SeO42− → SeO32− (selenate reductase). The conditions under which these function and their regulation depend on the organism. For example, in Escherichia coli, oxygen represses the synthesis of the other reductases, and under anaerobic conditions the reductases for fumarate, DMSO, and TMAO are repressed by nitrate. This does not apply to Wolinella succinogenes in which sulfur represses the synthesis of the more positive electron acceptors nitrate and fumarate (Lorenzen et al. 1993). The DMSO reductase from Escherichia coli (Weiner et al. 1988) has a broad substrate versatility, and is able to reduce a range of sulfoxides and N-oxides. Anaerobic sulfate reduction is not discussed here in detail.

OXYANIONS: NITRATE, SULFATE, CHLORATE, SELENATE, AND ARSENATE Nitrate and Related Compounds Nitrate can fulfill several functions: (a) assimilatory as a nitrogen source after reduction, (b) dissimilatory in reduction, and (c) respiration in which the reduction is coupled to the synthesis of ATP. The structural genes for assimilatory reduction are encoded by Nas genes and contain FAD, [4Fe–4S], and bis-molybdenum guanine dinucleotide (MGD); in dissimilation, the FAD is replaced by cytochrome b or c (Moreno-Vivián et al. 1999). The membrane-bound nitrate reductase that is involved in proton translocation and energy production is encoded by the genes NarGH, and the enzyme contains molybdopterin guanine dinucleotide, [Fe–S] clusters, and diheme cytochrome b556. The periplasmic nitrate reductase encoded by NapABC is primarily involved in dissimilation. The degradation of organic compounds with nitrate in the absence of oxygen—denitrification or nitrate dissimilation—has been known for a long time and has been used as a valuable diagnostic character in bacterial classification. The products are either dinitrogen or nitrous oxide, and the reaction is generally inhibited by oxygen, so that it occurs to a significant extent only under anoxic conditions. For example, although it has been reported that Thiosphaera pantotropha (Paracoccus denitrificans) is capable of both denitrification and nitrification under aerobic conditions (Robertson and Kuenen 1984), it has been shown that the suite of enzymes necessary for denitrification is not expressed constitutively. In addition, the rates of denitrification under aerobic conditions were very much slower than under anaerobic conditions (Moir et al. 1995). An important development is the appreciation of the role of anammox bacteria. These carry out the anaerobic reaction between NH4+ and NO2− in which nitrite is the electron donor with the production of N2 (Schmid et al. 2005). Several groups of bacteria have been implicated, and all of them belong to the phylum Planctomyces, although quite distinct organisms may be involved. They

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are widely distributed in habitats ranging from systems treating ammonia-rich waste to anaerobic marine environments (Tal et al. 2005). These organisms are unusual among anaerobic bacteria in many respects, and a number of procedures have been outlined to assess their contribution to nitrogen cycling, their enumeration, and their distribution. Denitrification involves the sequential formation of nitrite, nitric oxide, and nitrous oxide. Two aspects of nitric oxide have attracted attention: (a) chemical oxidation of biogenic nitric oxide to Nox in the context of increased ozone formation (Stohl et al. 1996) and (b) the physiological role in mammalian systems (Feldman et al. 1993; Stuehr et al. 2004), in parasitic infections (James 1995), and in the inhibition of bacterial respiration (Nagata et al. 1998). Nitric oxide may be produced microbiologically in widely different reactions such as a. Oxidation of l-arginine by a strain of Nocardia sp. produced nitric oxide and l-citrulline (Chen and Rosazza 1995). The enzyme (nitric oxide synthase) carries out two distinct reactions: (i) hydroxylation of l-arginine catalyzed by an enzyme analogous to cytochrome P450 and (ii) a one-electron oxidation of Nω-hydroxy-l-arginine to NO and l-citrulline. The mechanism has been reviewed (Stuehr et al. 2004). b. During the metabolism of glycerol trinitrate by Phanerochaete chrysosporium (Servent et al. 1991). Renewed interest has been focused on degradation of xenobiotics under anaerobic conditions in the presence of nitrate—possibly motivated by the extent of leaching of nitrate fertilizer from agricultural land into groundwater. In studies with such organisms, a clear distinction should be made between degradation of the substrate under three conditions that may or may not be biochemically equivalent: (a) aerobic conditions, (b) anaerobic conditions in the presence of nitrate, and (c) fermentation by strictly anaerobic conditions in the absence of any electron acceptor. Some examples are given to illustrate the diversity of degradations that have been observed with facultatively anaerobic organisms using nitrate as the electron acceptor: 1. The degradation of carbon tetrachloride to CO2 by a Pseudomonas sp. (Criddle et al. 1990), although a substantial part of the label was retained in nonvolatile water-soluble residues (Lewis and Crawford 1995). The nature of this was revealed by isolation of adducts with cysteine and N,N′-dimethylethylenediamine in which intermediates formally equivalent to COCl2 and CSCl2 were trapped, presumably formed by reaction of the substrate with water and a thiol, respectively. Further consideration of these reactions is given in Chapter 7, Part 3. 2. The nonstoichiometric production of trichloromethane from tetrachloromethane by Shewanella putrefaciens (Picardal et al. 1993). 3. The degradation of benzoate (Taylor and Heeb 1972; Williams and Evans 1975; Ziegler et al. 1987) and o-phthalate (Nozawa and Maruyama 1988; Afring and Taylor 1981). 4. The degradation of alkyl benzenes (Hutchins 1991; Evans et al. 1991a,b; Altenschmidt and Fuchs 1991). In these studies, some of the organisms referred to the genus Pseudomonas have been transferred to the genus Thauera (Anders et al. 1995). 5. The degradation of pristane in microcosms and in enrichment cultures (Bregnard et al. 1997). 6. The mineralization of cholesterol by an organism related to Rhodocyclus, Thauera, and Azoarcus (Harder and Probian 1997). It is also important to appreciate that an organism that can degrade a given substrate under conditions of nitrate dissimilation may not necessarily display this potential under aerobic conditions. For example, a strain of Pseudomonas sp. could be grown with vanillate under anaerobic conditions in the presence of nitrate but was unable to grow under aerobic conditions with vanillate.

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In contrast, cells grown anaerobically with nitrate and vanillate were able to oxidize vanillate under both aerobic and anaerobic conditions. The cells were also able to demethylate a much wider spectrum of aromatic methoxy compounds under anaerobic conditions than under aerobic conditions (Taylor 1983). Such subtleties should be clearly appreciated and taken into consideration in evaluating the degradative potential of comparable organisms under different physiological conditions. During degradation in the presence of nitrate this may be reduced, and it has been observed that nitro groups may be introduced into the substrate. Details of this have been discussed in Chapter 2. Sulfate and Related Compounds Sulfate can be either assimilated as sulfur source or dissimilated by anaerobic bacteria as an electron acceptor when it is reduced to sulfide. Both reactions involve the formation of AMP anhydrides (APS) catalyzed by ATP sulfurylase. In dissimilation, this is reduced successively to sulfite and sulfide. In assimilation, however, the ester is phosphorylated by ATP and an APS kinase to the anhydride, which after additional phosphorylation at C3′ (PAPS), undergoes further reduction by PAPS reductase encoded by the gene cysH (Bick et al. 2000). The assimilatory sulfite reductase from Escherichia coli is a complex enzyme containing two different polypeptides. The α-chain coded by the cysJ gene in Escherichia coli binds one of the flavins (FAD or FMN), while the β-chain coded by the cysI gene binds one [Fe4S4] cluster and one siroheme that is the site of sulfite reduction (Eschenbrenner et al. 1995). Siroheme is an iron tetrahydroporphrrin with eight carboxylate groups and two methyl groups on the reduced rings (Murphy et al. 1973). Although the metabolism of tetrathionate has not been established in detail, it presumably takes place via thiosulfate, sulfite, and sulfide. Tetrathionate, which is a component of some media for enrichment of salmonellas, is able to support the anaerobic growth of Salmonella sp. using glycerol or acetate, and S. enterica with ethanolamine or propan-1,2-diol when vitamin B12 is available (Price-Carter et al. 2001). Chlorate and Perchlorate Chlorate can serve as electron acceptor under anaerobic conditions (Thorell et al. 2003; Coates et al. 1999), and chlorate reductase has been found both in organisms such as Proteus mirabilis that can reduce chlorate but is unable to use to couple this to growth, and in true chlorate-respiring organisms. Chlorate has been shown to support the growth of an anaerobic community growing at the expense of acetate (Malmqvist et al. 1991), and a pure culture designated Ideonella dechloratans has been isolated (Malmqvist et al. 1994). A number of other organisms can use chlorate as electron acceptors during anaerobic growth and all of them can carry out dismutation of the intermediate chlorite. These include strain GR-1 that is able to carry out the sequential reduction of perchlorate to chlorate, and chloride at the expense of acetate (Rikken et al. 1996). Organisms capable of dissimilatory (per)chlorate reduction are both ubiquitous and diverse, and on the basis of 16S rRNA they are distinct from the previously described Ideonella dechloratans, and have been assigned to species of Dechloromonas (Achenbach et al. 2001). They can utilize C2–C5 carboxylic acids, lactate, fumarate, and malate as electron donors using chlorate as electron acceptor. Some of them can utilize both nitrate and chlorate as electron acceptors with acetate as electron donor (Coates et al. 1999), while Dechloromonas strain RCB, which is able to degrade benzene in the presence of nitrate, is also able to use perchlorate (Chakraborty et al. 2005). Pseudomonas chloritidismutans is clearly different from these organisms and is similar to P. stutzeri, except in the ability to use chlorate as electron acceptor. It is able to use chlorate, though not perchlorate or nitrate, as electron acceptors using acetate and propionate as sources of carbon (Wolterink et al. 2002). Although some isolates are capable of high rates of growth using both perchlorate and chlorate, some are able only to use the latter (Logan et al. 2001). Chlorate reductase has been characterized in strain GR-1 where it was found in the periplasm. It is oxygen-sensitive and contains molybdenum and [3Fe–4S] and [4Fe–4S] clusters (Kengen et al. 1999).

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A different cytoplasmic reductase has been characterized from Pseudomonas chloritidismutans (Wolterink et al. 2003). Chlorate reductase from I. dechloratans has been characterized (Thorell et al. 2003), and consists of three subunits, contains heme b, and resembles enzymes belonging to the molybdopterin DMSO reductase class II family. Perchlorate reductase from Dechloromonas agitata is different from that of I. dechloratans, although the pcrAB gene products are similar to α- and β-subunits of other reductases including nitrate, selenate, and chlorate. The pcrC gene product was similar to a c-type cytochrome, and the pcrD product to the molybdenum-containing proteins of the DMSO reductase family (Bender et al. 2005). The chlorite dismutase that generates oxygen from chlorite produced by reduction of chlorate has been purified and characterized from strain GR-1 (Van Ginkel et al. 1996) and from Ideonella dechloratans (Stenklo et al. 2001). The enzyme from Dechloromonas agitata has been suggested as a metabolic probe for chlorate reduction in the environment (Bender et al. 2002). Selenate and Arsenate Reductions in which energy is produced during reduction and supports growth of the cells should be carefully distinguished from the situation in which selenate is gratuitously reduced during aerobic growth (Maiers et al. 1988). For example, the membrane-bound selenate reductase from Enterobacter cloacae SLD1a-1 is not able to function as electron acceptor under anaerobic conditions (Ridley et al. 2006). Reductases also function in the detoxification of arsenate and selenate. The arsenate reductase gene ArsC both in Staphylococcus aureus (Ji and Silver 1992) and in Escherichia coli (Liu et al. 1995) is a determinant of arsenate resistance, and the arsenite that is produced is extruded from the cell by the ArsA–ArsB anion-translocating ATPase (Walmsley et al. 1999). Whereas the reductase from E. coli is coupled to the glutathione and glutaredoxin system, that from S. aureus requires thioredoxin and thioredoxin reductase. The arsenate reductase from Bacillus subtilis shows structural similarity at the active site to that of low-molecular-weight tyrosine phosphatase and this enabled a mechanism for its activity to be proposed (Bennett et al. 2001). Selenate and arsenate can serve as terminal electron acceptors, and strains of anaerobic bacteria have been isolated that are able to use them during growth with lactate. These include a strain designated SES-3 (Laverman et al. 1995), and Bacillus arsenicoselenatis and Bacillus selenitrificans (Blum et al. 1998). In these strains, utilization of acetate is unusual, but acetate is used specifically by Chrysiogenes arsenatis (Macy et al. 1996). The reductase from this organism contains Mo, Fe, and acid-labile S (Krafft and Macy 1998). A facultatively anaerobic organism Thauera selenatis is able to use both selenate and selenite as electron acceptors for anaerobic growth (Macy et al. 1993). The selenate reductase has been shown to contain molybdenum, iron, acid-labile sulfur, and cytochrome b. It is located in the periplasmic space, and is specific for selenate and nitrate, whereas neither chlorate nor sulfate was reduced at significant rates (Schröder et al. 1997). Reduction of selenite to Se0 can be accomplished by the periplasmic nitrite reductase that also occurs in this strain (DeMoll-Decker and Macy 1993). An organism assigned to Desulfotomaculum auripigmentum was able to grow with lactate and could use both arsenate and sulfate as terminal electron acceptors (Newman et al. 1997). Several strains of Desulfitobacterium can use arsenate, generally nitrate, and exceptionally selenate as electron acceptors during growth with lactate or pyruvate as electron donors (Niggemyer et al. 2001). A strain of Shewanella ANA-3 can couple the reduction of As(V) to As(III) to the oxidation of lactate to acetate (Saltikov et al. 2003). Although this strain contains the ars operon, which confers resistance to As(III), the gene arsB that encodes the efflux system and arsC for the reductase, although advantageous, are not obligatory. Genetic analysis showed the presence of an additional gene cluster that encoded a respiratory As(V) reductase. The gene arrA is predicted to encode a molybdenum cofactor, and arrB to encode an Fe–S cluster, and are predicted to be similar to those of the DMSO reductase family (Saltikov and Newman 2003). Therefore, although there are two available mechanisms for arsenate respiration, they differ in their expression: the ars system is

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expressed under both aerobic and anaerobic conditions, whereas the arr system is expressed only anaerobically and is repressed by oxygen or nitrate (Saltikov et al. 2005). A facultative autotroph (lithotroph) strain MLHE-1 was able to oxidize arsenite under anaerobic conditions to arsenate using nitrate as electron acceptor (Oremland et al. 2002).

METAL CATIONS AND OXYANIONS Many bacteria that ferment organic substrates are able to reduce Fe(III) to Fe(II) gratuitously, but these organisms are apparently unable to use the energy of this reduction for growth and couple this to the oxidation of organic substrates. There are, however, bacteria that can accomplish this, and some of them can also effectively use Mn(IV) and U(VI) (Lovley 1991) and, occasionally, vanadate. A wide range of organisms of different taxonomic affiliation can use Fe(III) as electron acceptor during growth with acetate. They include organisms currently assigned to the genera Geobacter, Pelobacter, and Desulfuromonas (Lovley et al. 1995; Lonergan et al. 1996; Coates et al. 2001), Shewanella putrefaciens (Myers and Nealson 1990), and Pantoea (Enterobacter) agglomerans (Francis et al. 2000). The anaerobe Desulfitobacter metallireducens is able to grow with lactate as electron donor using a range of electron acceptors including Fe(III) citrate, Mn(IV) oxide, 3-chloro-4-hydroxyphenylacetate, and humic substances (Finneran et al. 2002). Anaeromyxobacter dehalogenans is a chlororespiring organism that is able to couple the oxidation of acetate under anaerobic conditions to the reduction of ortho-halogenated phenols, nitrate, fumarate, and Fe(III) (He and Sanford 2003). V(V), Mn(IV), Fe(III), Tc(VII), and U(VI) There has been considerable interest in the remediation of sites contaminated with U(VI) and Tc(VII) by conversion of these soluble forms into less soluble reduced states that may be precipitated. In assessing results involving the reduction of U(VI) and Tc(VII), it is important to take into consideration (a) intermediate levels of reduction that may be critical and (b) the formation of complexes, for example, with carbonate that may determine the final products and their association with cells or particles. The discussion is assembled by organism, since some of these are able to reduce several oxidants. Shewanella spp. Strains of Shewanella putrefaciens (Alteromonas putrefaciens) are widely distributed in environmental samples and are generally considered as aerobic organisms with the capability of reducing thiosulfate to sulfide in complex media. They are also able, however, to grow anaerobically using Fe(III) as electron acceptor, and to oxidize formate, lactate, or pyruvate. These substrates cannot, however, be completely oxidized to CO2 since the acetate produced from the C2 compounds is not further metabolized. Mn[IV] may function analogously to Fe(III) (Lovley et al. 1989). The bioenergetics of the system has been examined in cells of another strain of this organism grown anaerobically with lactate using either fumarate or nitrate as electron acceptors. Respiration-linked proton translocation in response to Mn(IV), fumarate, or oxygen was clearly demonstrated (Myers and Nealson 1990). Levels of Fe(III) reductase, nitrate reductase, and nitrite reductase are elevated by growth under microaerophilic conditions, and the organism probably possesses three reductase systems, each of which apparently consists of low- or high-rate components (DiChristina 1992). Reductase activity is located on the outer membranes of Shewanella putrefaciens and has a requirement for cytochrome c (Myers and Myers 1997) and menaquinones (Saffarini et al. 2002). Another organism with hitherto unknown taxonomic affinity has been isolated (Caccavo et al. 1992), and is able to couple the oxidation of lactate to the reduction of Fe(III), Mn(IV), and U(VI). Under anaerobic conditions, Shewanella oneidensis MR1 can couple the reduction of V(V) to V(IV) to the oxidation of lactate, formate, and pyruvate (Carpentier et al. 2003). Respiration and growth can be carried out by reduction of vanadate: proton translocation across the cytoplasmic membrane takes place during the reduction and is abolished by CCCP, HOQNO, and antimycin (Carpentier et al. 2005).

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Under anaerobic conditions with H2 as electron donor—when growth did not take place—S. putrefaciens reduced Tc(VII) to Tc(IV) that was associated with the cell or with Tc(IV) complexes in the presence of carbonate (Wildung et al. 2000). Geobacter spp. A strictly anaerobic organism designated GS-15 and now assigned to the taxon Geobacter metallireducens (Lovley et al. 1993) is able to use Fe(III) as electron acceptor under anaerobic conditions for the oxidation of a number of substrates including acetate (Lovley and Lonergan 1990), and toluene and phenols (Champine and Goodwin 1991). This organism is also able to oxidize acetate by reduction of Mn(IV) and U(VI) (Lovley et al. 1993), and additional comments are given later in this chapter. Organisms belonging to the genus Geobacter are widely distributed in anaerobic environments in which Fe(III) occurs, and members of the Geobacteraceae were enriched at a site contaminated with U(VI) (Anderson et al. 2003). All species in the family are able to use Fe(III) as electron acceptor and acetate as electron donor, generally propionate and benzoate and, for G. metallireducens and G. grbiciae, additionally toluene using anthraquinone-2,6-disulfonate (AQDS) as an electron acceptor (Coates et al. 2001). The phylogenetically distinct Geovibrio ferrireducens strain PAL-1 is able to use a wide range of organic compounds as electron donors including acetate, propionate, succinate, and proline (Caccavo et al. 1996). It has been shown in Geobacter sulfurreducens that although U(VI) [UO22+] is reduced to insoluble U(IV), this occurs by initial reduction to U(V) followed by disproportionation (Renshaw et al. 2005). Therefore, it was suggested that caution be exercised in the use of microbial remediation for transuranic elements whose oxidation levels differ from those of uranium. There are other limitations to their application for bioremediation of uranium-contaminated sites: a. In long-term experiments, U(IV), which was initially produced by microbial reduction of U(VI), could be reoxidized putatively using residual Fe(III) or Mn(IV) even under anaerobic conditions (Wan et al. 2005). b. Uranium that is associated with sediments is resistant to reduction (Ortiz-Bernad et al. 2004a). It has been shown that in the presence of nitrate and an electron donor such as acetate, both U(VI) and Tc(VII) can be reduced (Istok et al. 2004). Important conclusions were that (a) whereas reduction of Tc(VII) occurred concurrently with nitrate reduction, reduction of U(VI) took place only after Fe(II) was detected and (b) reoxidation of U(IV) took place in the presence of high—though not with low—concentrations of nitrate. The reductase in Geobacter sulfurreducens is located in the outer membrane, and a soluble Fe(III) reductase has been characterized from cells grown anaerobically with acetate as electron donor and Fe(III) citrate or fumarate as electron acceptor (Kaufmann and Lovley 2001). The enzyme contained Fe, acid-labile S, and FAD. An extracellular c-type cytochrome is distributed in the membranes, the periplasm, and the medium, and functions as a reductase for electron transfer to insoluble iron hydroxides, sulfur, or manganese dioxide (Seeliger et al. 1998). The molecular mass of 9.6 kDa suggested its similarity to cytochromes from sulfate-reducing bacteria, which was consistent with the fact that it contained three hemes although there is conflicting evidence on the role of this component (Lloyd et al. 1999). In contrast, Fe(III) reduction in the hyperthermophilic archaeon Pyrobaculum islandicum and in Pelobacter carbinolicus (Lovley et al. 1995) seems not to involve cytochromes (Childers and Lovley 2001). Although the hyperthermophilic sulfate-reducing archaeon Archaeoglobus fulgidus contains an Fe(III)–EDTA reductase, this shows sequence similarity to the family of NAD(P)H:FMN oxidoreductases (Vadas et al. 1999). Cell suspensions of Geobacter sulfurreducens can couple the oxidation of hydrogen to the reduction of Tc(VII) to insoluble Tc(IV). An indirect mechanism involving Fe(II) was also observed, and was substantially increased in the presence of the redox mediator AQDS (Lloyd et al. 2000).

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Geobacter metallireducens is also able to grow by vanadate respiration supported by acetate, and it has been suggested that this could provide a new strategy for removing vanadate from groundwater (Ortiz-Bernad et al. 2004b). Other Organisms 1. Sulfate-reducing bacteria: A number of sulfate-reducing anaerobic bacteria can oxidize S0 to sulfate at the expense of Mn(IV) (Lovley and Phillips 1994), and a strain of Desulfovibrio desulfuricans was able to use hydrogen to couple the reduction of Tc(VII) to an insoluble, more reduced form that was precipitated on the periphery of the cells (Lloyd et al. 1999). Reduction of U(VI) complexes with aliphatic carboxylates was examined using Desulfovibrio desulfuricans: the acetate complex was more rapidly reduced than those of dicarboxylates such as malonate, oxalate, and citrate, and the complex with 4,5dihydroxybenzene-1,3-disulfonate was readily accessible (Ganesh et al. 1997). In the presence of H2, Desulfovibrio fructosovorans was able to reduce Tc(VII) to soluble Tc(V) or putatively to an insoluble precipitate of Tc(IV), and this was mediated by the Ni–Fe hydrogenase in the organism (De Luca et al. 2001). 2. Deinococcus radiodurans strain R1: Anaerobic cultures grown with lactate could reduce chromate, although reduction was increased by the presence of electron transfer to AQDS. Reduction of U(VI) and Tc(VII) could also be accomplished in the presence of the electron transfer agent (Fredrickson et al. 2000). 3. Pyrobaculum aerophilum: Reduction of Cr(VI) to Cr(III), U(VI) to insoluble U(IV), and Tc(VII) to insoluble Tc(IV or V) was accomplished with this hyperthermophilic organism in the presence of H2 (Kashefi and Lovley 2000). 4. Cellulomonas spp.: Under anaerobic conditions, several strains could carry out reduction of Cr(VI) and U(VI) at the expense of lactate (Sani et al. 2002). 5. Desulfitobacterium metallireducens: Growth occurred with lactate using Fe(III) and MnO2 as electron acceptors, and Cr(VI) is also reduced (Finneran et al. 2002). 6. Anaeromyxobacter dehalogenans: This organism that is able to use ortho-substituted phenols for chlororespiration is able to use various forms of Fe(III) (He and Sanford 2003). All these transformations illustrate important processes for the cycling of organic carbon in sediments where Fe(III) has been precipitated, and it seems likely that comparable geochemical cycles involving manganese (Lovley and Phillips 1988) will also achieve greater prominence (Lovley 1991; Nealson and Myers 1992). It has also been suggested that such organisms could be used for the immobilization of soluble U(VI) in wastewater containing both U(VI) and organic compounds by conversion into insoluble U(IV) (Lovley et al. 1991). The reductases that produce the insoluble Cr(III) from chromate are noted in Part 3 of this chapter.

HUMIC ACID AND ANTHRAQUINONE-2,6-DISULFONATE IN REDOX SYSTEMS Humic acid and the corresponding fulvic acid are complex polymers whose structures are incompletely resolved. It is accepted that the structure of humic acid contains oxygenated structures, including quinones that can function as electron acceptors, while reduced humic acid may carry out reductions. These have been observed both in the presence of bacteria that provide the electron mediator and in the absence of bacteria in abiotic reactions, for example, reductive dehalogenation of hexachloroethane and tetrachloromethane by anthrahydroquininone-2,6-disulfonate (Curtis and Reinhard 1994). Reductions using sulfide as electron donor have been noted in Chapter 1. Some experimental aspects are worth noting: a. Particularly in earlier experiments, a commercial preparation of humic acids, “Aldrich humic acid,” has been used and this differs significantly from the naturally occurring

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complex. Pure preparations are, however, now available from the International Humic Substances Society. b. Although AQDS has been used as a surrogate for the quinone structure of humic material, it differs from this in important respects including its water solubility and its redox potential. This applies equally to hydroquinone (H2AQDS). 1. Bacterial reduction of humic acid and AQDS: a. The frequency of humic-reducing bacteria has been assessed using AQDS as a surrogate for humic material and as electron acceptor, with acetate as electron donor. The organisms that were able to grow belonged to the Geobacteraceae and were able to reduce both humic acid and Fe(III) citrate, could therefore be significant in environments containing humic acid and Fe(III) (Coates et al. 1998). Geobacter metallireducens and G. grbiciae that were able to couple the oxidation of acetate or toluene to the reduction of Fe(III) could also use AQDS as electron acceptor (Coates et al. 2001). Desulfitobacterium metallireducens was able to grow anaerobically by coupling the reduction of AQDS to the oxidation of lactate to acetate and CO2 (Finneran et al. 2002). In addition, with lactate as electron donor, 3-chloro-4-hydroxyphenylacetae, tetrachloroethene, and trichloroethene could serve as electron acceptors. b. Fermentative bacteria including Propionibacterium freudenreichii were able to reduce humic acids and to produce relatively greater amounts of oxidized products. Specifically, during growth with lactate there was an increase in the formation of acetate at the expense of propionate. Reduction of humic acid was also observed with the fermentative Lactococcus lactis and Enterococcus cecorum (Benz et al. 1998). 2. Bacteria have been isolated using reduced anthraquinone-2,6-disulfonate (H2AQDS) as electron donor and nitrate as electron acceptor (Coates et al. 2002). The organisms belonged to the α-, β-, γ-, and δ-subdivision of the Proteobacteria, and were able to couple the oxidation of H2AQDS to the reduction of nitrate with acetate as the carbon source. In addition, a number of C2 and C3 substrates could be used including propionate, butyrate, fumarate, lactate, citrate, and pyruvate. 3. The microbial degradation of contaminants under anaerobic conditions using humic acids as electron acceptors has been demonstrated. These included the oxidations: (a) chloroethene and 1,2-dichloroethene to CO2 that was confirmed using 14C-labeled substrates (Bradley et al. 1998) and (b) toluene to CO2 with AQDS or humic acid as electron acceptors (Cervantes et al. 2001). The transformation of 1,3,5-trinitro-1,3,5-triazine was accomplished using Geobacter metallireducens and humic material with AQDS as electron shuttle (Kwon and Finneran 2006). 4. In a wider context, extracellular redox mediators have been implicated in a number of reductions. The specific role of reductive dehalogenation by porphyrins and corrins has been discussed in Chapter 1. a. The extracellular reductase from Geobacter sulfurreducens that is a c-type cytochrome can accept not only Fe(III) but also less readily AQDS and humic acids (Seeliger et al. 1998). b. The anaerobic reduction of azo dyes can be carried out by several bacteria including Sphingomonas xenophaga BN6 and Escherichia coli strain K-12 (Rau et al. 2002). During the degradation of naphthalene-2-sulfonate by Sphingomonas xenophaga BN6, 4-aminonaphtho-1,2-quinone and 4-ethanolaminonaphtho-1,2-quinone are produced and effectively enhance reduction by functioning as redox mediators (Keck et al. 2002). Analogously, the reduction of azo compounds by Escherichia coli is promoted by 2-hydroxy-1,4-naphthoquinone (lawsone) that is reduced by the O2-insensitive reductases NfsA and NfsB (Rau and Stolz 2003).

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c. The reduction of aromatic nitro compounds by the filtrate from a strain of Streptomyces sp. that is known to synthesize 2-amino-3-carboxy-5-hydroxybenzo-1,4-quinone (cinnaquinone) and the 6,6′-diquinone (dicinnaquinone) as secondary metabolites. The hydroquinone putatively carries out the reduction (Glaus et al. 1992).

DEHALORESPIRATION Dehalorespiration in which dehalogenation of chloroethenes is coupled to the synthesis of ATP has been demonstrated in a number of bacteria for several substrates and is discussed in Chapter 7, Part 3. Organisms include Sulfurospirillum (Dehalospirillum) multivorans, Dehalobacter restrictus, Desulfuromonas chloroethenica, and strains of Desulfitobacterium including Desulf. hafniense and Desulf. chlororespirans (references in Holliger et al. 1999) that can carry out reductive dechlorination of tetrachloroethene. In addition, Dehalococcoides ethenogenes is exceptional in being capable of reductively dehalogenating tetrachloroethene to ethene (Maymó-Gatell et al. 1999; Magnuson et al. 2000), while Desulfitobacterium hafniense (frappieri) has a wide spectrum of substrates that can be partially dechlorinated including chlorinated phenols, catechols, anilines, pentachloronitrobenzene, and pentachloropyridine (Dennie et al. 1998). The electron donors were generally H2 or pyruvate or methanol for Dehalococcoides ethenogenes. The chloroethene reductases contain cobalamin and Fe–S clusters, and the range of substrates for trichloroethene reductase from Dehalococcoides ethenogenes includes tribromoethene and bromoethene (Magnuson et al. 2000). In addition, the TCE reductive dehalogenase from this organism is able to debrominate substrates containing C2, C3, and C4 or C5 carbon atoms albeit with decreasing facility: illustrative reactions are the debromination of tribromoethene to dibromo-ethenes, vinyl bromide, and ethene. In a wider context, attention is drawn to debrominations carried out by methanogenic bacteria—ethene from 1,2-dibromoethane and ethyne from 1,2-dibromoethene (Belay and Daniels 1987). Details of these dehalogenations have emerged from studies with methanogens. The formation of ethene from 1,2-dichloroethane with hydrogen as electron donor has been demonstrated in cell extracts of Methanobacterium thermoautotrophicum DH, and in Methanosarcina barkeri it has been shown to involve cobalamin and F430 using Ti(III) as reductant (Holliger et al. 1992).

ALKANE SULFONATES AS TERMINAL ELECTRON ACCEPTORS A few aliphatic sulfonates such as 2-hydroxyethylsulfonate, alanine-3-sulfonate, and acetaldehyde2-sulfonate are able to serve as sulfur sources and electron acceptors during anaerobic growth of some sulfate-reducing bacteria when lactate is supplied as the carbon source (Lie et al. 1996). Several sulfite-reducing species of Desulfitobacterium are able to use 2-hydroxyethanesulfonate as terminal electron acceptor producing acetate and sulfide (Lie et al. 1999). It is relevant to note that species of Desulfitobacterium are also able to use chlorinated organic compounds and arsenate as terminal electron acceptors. Conversely, a strain of Rhodopseudomonas palustris isolated by enrichment with taurine was able to use this as a source of sulfur and nitrogen, and as electron source during photoautotrophic growth in the presence of CO2 (Novak et al. 2004).

NITROALKANES AS ELECTRON ACCEPTOR An anaerobic organism isolated from the rumen by enrichment for the metabolism of the toxic aglycone of miserotoxin (3-nitro-1-propyl-β-d-glucopyramnoside) was assigned to Denitrobacterium detoxificans (Anderson et al. 2000). It was able to use a number of nitroalkanes including 3-nitropropanol, 3-nitropropionate, nitroethane, and 2-nitrobutane as electron acceptors during growth with lactate that was converted into acetate.

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Picardal FW, RG Arnold, H Couch, AM Little, ME Smith (1993) Involvement of cytochromes in the anaerobic biotransformation of tetrachloromethane by Shewanella putrefaciens 200. Appl Environ Microbiol 59: 3763–3770. Price-Carter M, J Tingey, TA Bobik, JR Roth (2001) The alternative electron acceptor tetrathionate supports B12-dependent anaerobic growth of Salmonella enterica serovar typhimurium on ethanolamine or 1,2propandiol. J Bacteriol 183: 2463–2475. Rau J, A Stolz (2003) Oxygen-insensitive nitroreductases NfsA and NfsB of Escherichia coli function under anaerobic conditions as lawsone-dependent azo reductases. Appl Environ Microbiol 69: 3448–3455. Rau J, H-J Knackmuss, A Stolz (2002) Effects of different quinoid redox mediators on the anaerobic reduction of azo dyes by bacteria. Environ Sci Technol 36: 1497–1504. Renshaw JC, LJC Butchins, FR Livens, I May, JR Lloyd (2005) Bioreduction of uranium: environmental implications of a pentavalent intermediate. Environ Sci Technol 39: 5657–5660. Ridley H, CA Watts, DJ Richardson, CS Butler (2006) Resolution of distinct membrane-bound enzymes from Enterobacter cloacae SKLD1a-1 that are responsible for selective reduction of nitrate and selenate anions. Appl Environ Microbiol 72: 5173–5180. Rikken GB, AGM Croon, CG van Ginkel (1996) Transformation of perchlorate into chloride by a newly isolated bacterium: reduction and dismutation. Appl Microbiol Biotechnol 45: 420–426. Robertson LA, JG Kuenen (1984) Aerobic denitrification: a controversy revived. Arch Microbiol 139: 351–354. Saffarini DA, SL Blumerman, KJ Mansoorabadi (2002) Role of menaquinones in Fe(III) reduction by membrane fractions of Shewanella putrefaciens. J Bacteriol 184: 846–848. Saltikov CW, A Cifuentes, K Venkateswaran, DK Newman (2003) The ars detoxification system is advantageous but not required for As(V) respiration by the genetically tractable Shewanella species strain ANA-3. Appl Environ Microbiol 69: 2800–2809. Saltikov CW, DK Newman (2003) Genetic identification of a respiratory arsenate reductase. Proc Natl Acad Sci USA 100: 10983–10988. Saltikov CW, RA Wildman, DK Newman (2005) Expression dynamics of arsenic respiration and detoxification in Shewanella sp. strain ANA-3. J Bacteriol 187: 7390–7396. Sani RK, BM Peyton, WA Smith, WA Apekl, JN Petersen (2002) Dissimilatory reduction of Cr(VI), Fe(III), and U(VI) by Cellulomonas isolates. Appl Microbiol Biotechnol 60: 192–199. Schmid MC et al. (2005) Biomarkers for in situ detection of anaerobic ammonium-oxidizing (Anammox) bacteria. Appl Environ Microbiol 71: 1677–1684. Schröder I, S Rech, T Krafft, JM Macey (1997) Purification and characterization of the selenate reductase from Thauera selenatis. J Biol Chem 272: 23765–23768. Seeliger S, R Cord-Ruwisch B Schink (1998) A periplasmic and extracellular c-type cytochrome of Geobacter sulfurreducens acts as a ferric iron reductase and as an electron carried to other acceptors or to partner bacteria. J Bacteriol 180: 3686–3691. Servent D, C Ducrorq, Y Henry, A Guissani, M Lenfant (1991) Nitroglycerin metabolism by Phanerochaete chrysosporium: evidence for nitric oxide and nitrite formation. Biochim Biophys Acta 1074: 320–325. Stenklo K, HD Thorell, H Bergius, R Aasa, T Nilsson (2001) Chlorite dismutase from Ideonella dechloratans. J Biol Inorg Chem 6: 601–607. Stohl A, E Williams, G Wotawa, H Kromp-Kolb (1996) A European inventory of soil nitric oxide emissions and the effect of these emissions on the photochemical formation of ozone. Atmos Environ 30: 3741–3755. Stuehr DJ, J Santolini, Z-Q Wang, C-C Wei, S Adak (2004) Update on mechanism and catalytic regulation in the NO synthases. J Biol Chem 279: 36167–36170. Tal Y, JEM Watts, HJ Svchreier (2005) Anaerobic ammonia-oxidizing bacteria and related activity in Baltimore inner harbor sediment. Appl Environ Microbiol 71: 1816–1821. Taylor BF (1983) Aerobic and anaerobic catabolism of vanillic acid and some other methoxy-aromatic compounds by Pseudomonas sp. strain PN-1. Appl Environ Microbiol 46: 1286–1292. Taylor BF, MJ Heeb (1972) The anaerobic degradation of aromatic compounds by a denitrifying bacterium. Radioisotope and mutant studies. Arch Microbiol 83: 165–171. Thorell HD, K Stenklo, J Karlsson, T Nilsson (2003) A gene cluster for chlorate metabolism in Ideonella dechloratans. Appl Environ Microbiol 69: 5585–5592. Vadas A, HG Monbouquette, E Johnson, I Schröder (1999) Identification and characterization of a novel ferric reductase from the hyperthermophilic archaeon Archaeoglobus fulgidus. J Biol Chem 274: 36715–36721.

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PART 3 REDUCTASES AND RELATED ENZYMES Reduction is an important reaction under both aerobic and anaerobic conditions. Reductases mediate a wide variety of reactions that are summarized briefly here, and have been discussed in detail in Part 2 dealing with electron acceptors and further in Part 5 with metalloenzymes. The reductases that are components of the aromatic dioxygenases and that are involved in the aerobic bacterial degradation of aromatic hydrocarbons are noted parenthetically in Chapter 8, Parts 1 and 2.

REDUCTION OF NITROARENES Nitroarene reductases catalyze the reduction of nitroarenes. They are flavoprotein (FMN) enzymes (Blehert et al. 1999) and are of two types: 1. Type I are O2-insensitive six-electron reductases that catalyze the sequential reduction of nitroarenes to nitroso, hydroxylamino, and amino arenes. They are encoded in E. coli by NfsA for the major enzyme and NfsB for the minor (Rau and Stolz 2003), and are also important in establishing resistance to nitrofuran drugs (Koziarz et al. 1998). 2. Type II are O2-sensitive two-electron reductases that catalyze the reduction to ArNO2 radicals that react with O2 to produce superoxide (Bryant and DeLuca 1991; Bryant et al. 1991).

NITRATE ESTER REDUCTASE The degradation of aliphatic nitrate esters to nitrite and the corresponding alkanol takes place by a reductive rather than a hydrolytic mechanism. The reduction of glycerol trinitrate by Agrobacterium radiobacter (Snape et al. 1997) and of pentaerythritol tetranitrate by Enterobacter cloacae strain PB2 (French et al. 1996) have been described, and of flavoprotein reductases from Pseudomonas fluorescens strain I-C (Blehert et al. 1999) have been characterized. They belong to a group of flavoproteins that are related to “old yellow enzyme” with varying substrate specificities, and can reduce pentaerythritol tetranitrate, 1-nitrocyclohexene, cyclohex-2-enone, N-ethylmaleamide, and morphinone (Williams et al. 2004). It has been shown that hydrogen transfer in the reaction with NADH in pentaerythritol tetranitrate and morphinone reductase involves hydrogen tunneling in the reaction between the reduced flavin and the substrate (Basran et al. 2003).

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ENONES (α, β-UNSATURATED KETONES) AND RELATED REDUCTASES A group of flavoenzyme reductases mediate the reduction both nitrate esters and C=C double bonds generally activated with carbonyl or nitro functions. The structures of the substrates vary widely (references in Faber 1997). Aerobic Conditions 1. The reduction of cyclohex-2-ene-1-one to cyclohexanone is highly specific in Pseudomonas syringae (Rohde et al. 1999), whereas the reductases from Saccharomyces cerevisiae are able to accept a wide range of activated aliphatic enones (Wanner and Tressel 1998). The related “old yellow enzyme” and pentaerythritol tetranitrate reductase can carry out reduction of androsta-∆1,4-3,17-dione to androsta-∆4 -3,17-dione (Vaz et al. 1995), and prednisone to pregna-∆4-17α,20-diol-3,11,20-trione (Barna et al. 2001). Old yellow enzyme is also able to bring about the unusual dismutations of several conjugated cyclohexenones and cyclodecenones (Vaz et al. 1995) that are discussed in Chapter 7, part 2. 2. A reductase in Escherichia coli can reduce N-ethylmaleamide to N-ethylsuccinimide (Miura et al. 1997). 3. Morphinone can be reduced by Pseudomonas putida M10 to hydromorphone using an enzyme of which one of the subunits contains FMN (French and Bruce 1994). 4. The reductase from Burkholderia sp. strain WS produces (S)-2-chloropropionate from 2-chloroacrylate (Kurata et al. 2005). 5. An enzyme that catalyzes the reduction of ∆1-piperidein-2-carboxylate to piperidine-2-carboxylate (l-pipecolate) in the catabolism of d-lysine by Pseudomonas putida ATCC12633 is an NADPH-dependent representative of a large family of reductases that are distributed among bacteria and archaea (Muramatsu et al. 2005). It also catalyzes the reduction of ∆1-pyrrolidine-2-carboxylate to l-proline. Anaerobic Conditions 1. The reductases from Clostridium tyrobutyricum and Cl. kluyveri can reduce both aliphatic enoates and cinnamates to the dihydro compounds (Bühler and Simon 1982). The reductase in Cl. tyrobutyricum contains iron, labile sulfur, and both FAD and FMN (Kuno et al. 1985), while electron paramagnetic resonance (EPR) measurements suggest the presence of a semiquinone radical and Mössbauer spectra the presence of a [4Fe–4S] cluster (Caldeira et al. 1996). The amino acid sequences of these enzymes show similarity to a group of oxidoreductases and dehydrogenases, including the dehydrogenase for dimethylamine in Hyphomicrobium X and trimethylamine in Methylotrophus methylophilus (Rohdich et al. 2001). 2. Important reductions are involved in the anaerobic transformations of steroid and flavanoids: a. Reduction of a steroid 3-keto-4,6-diene to the 3-keto compound involved a complex sequence of reactions in a Clostridium sp. (Wells and Hylemon 2000), which is discussed in Chapter 7, Part 2. b. The reductions of the B-ring of flavanoids have been observed in Clostridium orbiscindens (Schoefer et al. 2003) and in Eubacterium ramulus (Braune et al. 2001), and these are noted again in Chapter 10, Part 2.

AZO REDUCTASE Reduction of the azo group in dyes and colorants is the key reaction in their decolorization. Although an azo reductase mediates the reduction of azo groups to amines, the metabolic situation is quite complex. The enzyme is apparently synthesized under both aerobic (Blümel et al. 2002) and

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anaerobic conditions (Rau et al. 2002; Rau and Stolz 2003), and the anaerobic reduction is facilitated by the presence of quinonoid mediators. The enzyme from Staphylococcus aureus is tetrameric with four noncovalently bound FMN and requires NADPH as reductant for activity (Chen et al. 2005).

ALDEHYDE OXIDOREDUCTASES There are diverse reactions that are mediated by aldehyde oxidoreductases: 1. The aerobic reduction of aryl and alkyl carboxylates to the corresponding aldehydes. The reaction involves formation of an acyl-AMP intermediate by reaction of the carboxylic acid with ATP; NADPH then reduces this to the aldehyde (Li and Rosazza 1998; He et al. 2004). The oxidoreductase from Nocardia sp. is able to accept a range of substituted benzoic acids, naphthoic acids, and a few heterocyclic carboxylic acids (Li and Rosazza 1997). 2. The phenylacetaldehyde reductase involved in the degradation of styrene is also able to accept long-chain aliphatic aldehydes and ketones, and halogenated acetophenones (Itoh et al. 1997). 3. There are several 2-ketoglutarate anaerobic oxidoreductases, for example, in Thauera aromatica (Dörner and Boll 2002) and Azoarcus evansii (Ebenau-Jehle et al. 2003). Their role in the metabolism of arene carboxylates is discussed in Chapter 8, Part 3. 4. The oxidoreductase from Pseudomonas diminuta strain 7 that carries out hydroxylation of isoquinoline at C2 is a molybdenum enzyme containing [Fe–S] centers, which is comparable to the aldehyde oxidoreductase from Desulfovibrio gigas (Lehmann et al. 1994).

F420-DEPENDENT (DESAZATETRAHYDROFOLATE) REDUCTASES These reductases play a key role both in methanogenesis and in the degradation of phenols that carry several nitro groups, which is discussed further in Chapter 9, Part 5. Although these reductases are typically found in methanogens, they have been encountered in a number of other bacteria and archaea: 1. Methanogenesis from CO2 via methenyl H4MPT, methylene H4MPT, and methyl H4MPT using coenzyme F420H2 (5,6,7,8-tetrahydrofolate). 2. Degradation of 2,4,6-trinitrophenol and 2,4-dinitrophenol involving a Meisenheimer hydride complex (Vorbeck et al. 1998; Behrend and Heesche-Wagner 1999; Ebert et al. 1999; Hofmann et al. 2004; Heiss et al. 2002). 3. The sulfite reductase from the hyperthermophilic methanogen Methanocaldococcus jannashii is able to reduce the otherwise toxic sulfite to sulfide that is required for growth. In contrast to most organisms that use nicotinamides and cytochromes as electron carriers, this organism uses a coenzyme F420 –dependent reductase (Johnson and Mukhopadhyay 2005). 4. Glucose-6-phosphate is normally dehydrogenated to 6-phosphogluconate using an NADPdependent dehydrogenase. In Mycobacterium smegmatis strain mc26, however, the reaction is mediated by coenzyme F420, and neither NAD nor NADP was effective (Purwantini and Daniels 1996).

OTHER CARBOCYCLIC AROMATIC COMPOUNDS Reduction of Carbocyclic Rings Reductases play a cardinal role in the anaerobic degradation of diverse aromatic compounds, for example, (a) the reduction of benzoyl-CoA that is a prelude to the fission of the ring (Harwood et al. 1999) (Chapter 8, Part 3) and (b) the degradation of phloroglucinol via dihydrophloroglucinol to acetate and butyrate (Haddock and Ferry 1989, 1993) (Chapter 8, Part 4).

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Dehydroxylation The anaerobic degradation of some hydroxybenzoates and phenols involves reductive removal of the phenolic hydroxyl group. The enzyme that dehydroxylates 4-hydroxybenzoyl-CoA in Thauera aromatica is a molybdenum–flavin–iron–sulfur protein (Breese and Fuchs 1998), and is similar to the enzyme from the nonsulfur phototroph Rhodopseudomonas palustris that carries out the same reaction (Gibson et al. 1997). Reduction of Azaarene Rings The aerobic degradation of several azaarenes involves reduction of the rings at some stage, and are discussed in Chapter 10, Part 1. Illustrative examples include the degradation of pyridines (3-alkylpyridine, pyridoxal) and pyrimidines (catalyzed by dihydropyrimidine dehydrogenases). Reductions are involved in both the aerobic and the anaerobic degradation of uracil and orotic acid.

METAL CATIONS AND OXYANIONS Although some metal cations and oxyanions can serve as electron acceptors for growth under anaerobic conditions, reduction may also take place gratuitously, and these reductases have been characterized in a number of organisms. Reduction has been implicated in resistance to metal cations and metalloid oxyanions, which are discussed in Part 4 of this chapter. 1. Soluble reductases with a temperature optimum of 80°C have been described from (a) Pseudomonas putida that reduces chromate to insoluble Cr(III) (Park et al. 2000) and (b) Archaeoglobus fulgidus that can reduce Fe(III)–EDTA (Vadas et al. 1999). A membrane-bound chromate reductase has been purified from Enterobacter cloacae (Wang et al. 1990). 2. Dimeric flavoprotein chromate reductases have been purified from Pseudomonas putida (ChrR) and Escherichia coli (YieF). The former produces a semiquinone and transiently reactive oxygen species, whereas the latter is an obligate four-electron reductant. Oneelectron reduction of Cr(VI) to Cr(V) has, however, been observed as an intermediate in the reduction by the NAD(P)H-dependent reductase of Pseudomonas ambigua strain G-1 (Suzuki et al. 1992). 3. It has been shown that the nitroreductases from Escherichia coli and Vibrio harveyi are homologous to the chromate reductase from Pseudomonas ambigua (Kwak et al. 2003). 4. Reductase activity is located on the outer membranes of Shewanella putrefaciens and has a requirement for cytochrome c (Myers and Myers 1997) and menaquinones (Saffarini et al. 2002). 5. The reductase in Geobacter sulfurreducens is located in the outer membrane and a soluble Fe(III) reductase has been characterized from cells grown anaerobically with acetate as electron donor and Fe(III) citrate or fumarate as electron acceptor (Kaufmann and Lovley 2001). The enzyme contained Fe, acid-labile S, and FAD. An extracellular c-type cytochrome is distributed in the membranes, the periplasm, and the medium, and functions as a reductase for electron transfer to insoluble iron hydroxides, sulfur, or manganese dioxide (Seeliger et al. 1998). 6. The selenate reductase from Enterobacter cloacae SLD1a-1 functions only under aerobic conditions, and is not able to serve as an electron acceptor for anaerobic growth, in contrast to the periplasmic enzyme from Thauera selenatis (Schröder et al. 1997). In E. cloacae there are separate nitrate and selenate reductases, both of which are membrane-bound. The selenate reductase is able to reduce chlorate and bromate though not nitrate, contains Mo, heme and nonheme iron, and consists of three subunits in an α3β3γ3 configuration.

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Hofmann KW, H-J Knackmuss, G Heiss (2004) Nitrite elimination and hydrolytic ring cleavage in 2,4,6-trinitrophenol (picric acid) degradation. Appl Environ Microbiol 70: 2854–2860. Itoh N, R Morihama, J Wang, K Okada, N Mizuguchi (1997) Purification and characterization of phenylacetaldehyde reductase from a styrene-assimilating Corynebacterium strain, ST-10. Appl Environ Microbiol 63: 3783–3788. Johnson EF, B Mukhopadhyay (2005) A new type of sulfite reductase, a novel coenzyme F420-dependent enzyme, from the methanarchaeon Methanocaldocccus jannaschii. J Biol Chem 280: 38776–38786. Kaufmann F, DR Lovley (2001) Isolation and characterization of a soluble NADPH-dependent Fe(III) reductase from Geobacter sulfurreducens. J Bacteriol 183: 4468–4476. Koziarz JWP, J Veall, N Sandhu, P Kumar, B Hoecher, IB Lambert (1998) Oxygen-insensitive nitroreductases: analysis of the roles of nfsA and nfsB in development of resistance to 5-nitrofuran derivatives in Escherichia coli. J Bacteriol 180: 5529–5539. Kuno S, A Bacher, H Simon (1985) Structure of enoate reductase from a Clostridium tyrobutyricum (C. spec. La1). Biol Chem Hoppe-Seyler 366: 463–472. Kurata A, T Kurihara, H Kamachi, N Esaki (2005) 2-haloacrylate reductase: a novel enzyme of the mediumchain dehydrogenase/reductase superfamily that catalyzes the reduction of carbon–carbon double bond of unsaturated organohalogen compounds. J Biol Chem 280: 20286–20291. Kwak YH, DS Lee, HB Kim (2003) Vibrio harveyi nitroreductase is also a chromate reductase. Appl Environ Microbiol 69: 4390–4395. Lehmann M, B Tshisuaka, S Fetzner, P Röger, F Lingens (1994) Purification and characterization of isoquinoline 1-oxidoreductase from Pseudomonas diminuta 7, a molybdenum-containing hydroxylase. J Biol Chem 269: 11254–11260. Li T, JPN Rosazza (1997) Purification, characterization, and properties of an aryl aldehyde oxidoreductase from Nocardia sp. strain NRRL 5646. J Bacteriol 179: 3482–3487. Li T, JPN Rosazza (1998) NMR identification of an acyl-adeylate intermediate in the aryl-aldehyde oxidoreductase catalyzsed reaction. J Biol Chem 273: 34230–34233. Miura K, Y Tomioka, H Suzuki, M Yonezawa, T Hishinuma, M Mizugaki (1997) Molecular cloning of the nemA gene encoding N-ethylmaleimide reductase from Escherichia coli. Biol Pharm Bull 20: 110–112. Muramatsu H, H Mihara, R Kakutani, M Yasuda, M Ueda, T Kurihara, N Esaki (2005) The putative malate/ lactate dehydrogenase from Pseudomonas putida is an NADPH-dependent ∆ 1-piperideine-2-carboxylate/∆1-pyrroline-2-carboxylate reductase involved in the catabolism of L-lysine and D-proline. J Biol Chem 280: 5329–5335. Myers CR, JD Myers (1997) Cloning and sequence of cymA, a gene encoding a tetraheme cytochrome c required for reduction of iron (III), fumarate, and nitrate by Shewanella putrefaciens MR-1. J Bacteriol 179: 1143–1152. Park CH, M Keyhan, B Wielinga, S Fendorf, A Matin (2000) Purification to homogeneity and characterization of a novel Pseudomonas putida chromate reductase. Appl Environ Microbiol 66: 1788–1795. Purwantini E, L Daniels (1996) Purification of a novel coenzyme F420 -dependent glucose-6-phosphate dehydrogenase from Mycobacterium smegmatis. J Bacteriol 178: 2861–2866. Rau J, A Stolz (2003) Oxygen-insensitive nitroreductases NfsA and NfsB of Escherichia coli function under anaerobic conditions as lawsone-dependent azo reductases. Appl Environ Microbiol 69: 3448–3455. Rau J, H-J Knackmuss, A Stolz (2002) Effects of different quinoid redox mediators on the anaerobic reduction of azo dyes by bacteria. Environ Sci Technol 36: 1497–1504. Ridley H, CA Watts, DJ Richardson, CS Butler (2006) Resolution of distinct membrane-bound enzymes from Enterobacter cloacae SKLD1a-1 that are responsible for selective reduction of nitrate and selenate anions. Appl Environ Microbiol 72: 5173–5180. Rohdich F, A Wiese, R Feiucht, H Simnon, A Bacher (2001) Enoate reductases of Clostridia. Cloning, sequencing, and expression. J Biol Chem 276: 5779–5787. Rohde BH, R Schmid, MS Ullrich (1999) Thermoregulated expression and characterization of an NAD(P)Hdependent 2-cyclohexen-1-one reductase in the plant pathogenic bacterium Pseudomonas syringae pv. glycinea J Bacteriol 181: 814–822. Saffarini DA, SL Blumerman, KJ Mansoorabadi (2002) Role of menaquinones in Fe(III) reduction by membrane fractions of Shewanella putrefaciens. J Bacteriol 184: 846–848. Schoefer L, R Mohan, A Schwiertz, A Braune, M Blaut (2003) Anaerobic degradation of flavonoids by Clostridium orbiscindens. Appl Environ Microbiol 69: 5849–5854. Schröder I, S Rech, T Krafft, JM Macey (1997) Purification and characterization of the selenate reductase from Thauera selenatis. J Biol Chem 272: 23765–23768.

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Seeliger S, R Cord-Ruwisch, B Schink (1998) A periplasmic and extracellular c-type cytochrome of Geobacter sulfurreducens acts as a ferric iron reductase and as an electron carried to other acceptors or to partner bacteria. J Bacteriol 180: 3686–3691. Snape JR, NA Walkley, AP Morby, S Nicklin, GF White (1997) Purification, properties, and sequence of glycerol trinitrate reductase from Agrobacterium radiobacter. J Bacteriol 179: 7796–7802. Suzuki T, N Miyata, H Horitsu, K Kawai, K Takamizawa, Y Tai, M Okazaki (1992) NAD(P)H-dependent chromium (VI) reductase of Pseudomonas ambigua G-1: a Cr(V) intermediate is formed during the reduction of Cr(VI) to Cr(III). J Bacteriol 174: 5340–5345. Vaz ADN, S Chakraborty, V Massey (1995) Old yellow enzyme: aromatization of cyclic enones and the mechanism of a novel dismutation reaction. Biochemistry 34: 4246–4256. Vorbeck C, H Lenke, P Fischer, JC Spain, H-J Knackmuss (1998) Initial reductive reactions in aerobic microbial metabolism of 2,4,6-trinitrotoluene. Appl Environ Microbiol 64: 246–252. Wang P, T Mori, K Toda, H Ohtake (1990) Membrane-associated chromate reductase activity from Enterobacter cloacae. J Bacteriol 172: 1670–1672. Wanner P, R Tressel (1998) Purification and characterization of two enone reductases from Saccharomyces cerevisiae. Eur J Biochem 255: 271–278. Wells JE, PB Hylemon (2000) Identification and characterization of a bile acid 7α-dehydroxylation operon in Clostridium sp. strain TO-931, a highly active 7α-dehydroxylating strain isolated from human feces. Appl Environ Microbiol 66: 1107–1113. Williams RE, DA Rathbone, NS Scrutton, NC Bruce (2004) Biotransformation of explosives by the Old Yellow Enzyme family of flavoproteins. Appl Environ Microbiol 70: 3566–3574.

PART 4 MICROBIAL REACTIONS TO CHEMICAL STRESS INTRODUCTION Microbial cells may be subjected to stress from a number of sources in their environment. These include antibiotics, metal cations and metalloid oxyanions, aromatic hydrocarbons, chlorophenols, and high acidity. Tolerance and resistance can be mediated by a number of mechanisms including modification of the toxicant and the operation of effective efflux systems. Hydrocarbon Tolerance Gram-Negative Bacteria A number of pseudomonads have been shown to tolerate high concentrations of aromatic hydrocarbons. For example, a strain of P. putida DOT-T1 was able to grow in the presence of 90% (v/v) toluene, and with a range not only of aromatic hydrocarbons including toluene, styrene, ethylbenzene, but also long-chain n-alkanes (Ramos et al. 1995). Advantage of this tolerance was taken by using the hydroxylase (tmoABCDEF) from this strain to produce 4-hydroxytoluene that could then undergo successive side-chain oxidation to 4-hydroxybenzoate (Ramos-González et al. 2003). Bacterial degradation of highly lipophilic aromatic compounds including phenol and hydrocarbons such as toluene necessitates their resistance to potentially adverse effects, and a number of mechanisms whereby toxicity is overcome have been considered (Sikkema et al. 1995). This may be accomplished by alteration of the structure of their lipid membranes although none of the mechanisms that have been proposed are entirely conclusive or exclusive. Associated Alterations in Fatty Acid and Lipid Structure It is important to distinguish short-term from long-term responses. In Pseudomonas putida DOT-T1, which is a solvent-tolerant strain, the short-term response has been associated with transformation of cis-9,10-methylene hexadecanoic acid to 9-cis-hexadecenoate and then to the trans isomer (Ramos et al. 1997). Pseudomonas putida strain S12 is able to grow with concentrations of toluene up to 50% and growth with acetate in the presence of 1% toluene was accompanied by the replacement of cis unsaturated fatty acids with the trans isomers. This was maintained during several generations even in the absence of toluene (Weber et al. 1994). Enzymes that bring about the cis/trans

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isomerization of unsaturated fatty acids (Cti) have been described in Pseudomonas oleovorans GPo12 that has been cured of the plasmid determining the degradation of octanol (Pedrotta and Witholt 1999). Although greater activity could be shown in the presence of crude membrane preparations from Pseudomonas putida DOT-T1 in the presence of alcohols, the gene cti for this enzyme was present even in solvent-sensitive strains (Junker and Ramos 1999). Therefore, although the presence of this enzyme is important in alleviating short-term damage to solvents, it cannot be the primary cause of solvent resistance in this strain. It has been pointed out that details of how the organisms are harvested and the samples prepared could jeopardize the results, and that the cis–trans isomerization is competitive with the formation of cyclopropane fatty acids from the cis isomers (Härtig et al. 2005). The synthesis of classes of phospholipid (phosphatidylethanolamine, phosphatidylglycerol, cardiolipin, phosphatidic acid, and phosphatidylserine) was examined in two strains of Pseudomonas putida, neither of which could degrade o-xylene and one of which was tolerant to o-xylene. For all strains, there were alterations in the concentration of total phospholipid, with a decrease in the sensitive strain and an increase in the tolerant strain. The fatty acid composition of the lipids was also altered in the tolerant strain, with a general decrease in cis unsaturated and increase in trans unsaturated fatty acids (this has already been noted). There was a higher rate of phospholipid biosynthesis in the tolerant strain and this suggests that the tolerant strain was able to repair membranes damaged by o-xylene more effectively than the sensitive strain (Pinkart and White 1997). In a wider context, alterations in membrane structure are generally associated with environmental stress. For example, it has been shown in the archaeon Methanococcoides burtonii that the degree of unsaturated lipids in cells grown at 4°C was higher than in cells grown at 23°C, and that the degree of unsaturation is specific for each class of phospholipid (Nichols et al. 2004). Efflux Pumps Gram-Negative Bacteria

An alternative—and possibly complementary—mechanism that is highly attractive is the existence of an efflux pump. Evidence has been provided for an energy-dependent export system in a toluene-resistant strain of Pseudomonas putida S-12 (Isken and de Bont 1996). The genes for this system have been cloned and their nucleotide sequence determined (Kieboom et al. 1998). Proteins coded by the three genes srpA, srpB, and srpC have extensive similarity to those for proton-dependent multidrug efflux systems, which are discussed in the section dealing with antibiotic resistance. In addition, the genes could be transferred to another strain of P. putida with the development of solvent resistance. Consistent with the role of the MexA-MexB-OprM operon that is expressed in Pseudomonas aeruginosa under normal growth conditions and that mediates antibiotic resistance, these genes also contribute to tolerance to n-hexane and p-xylene—though not toluene—in this strain (Li et al. 1998). As with antibiotic resistance, tolerance is abolished in the presence of the protonophore CCCP. In the toluene-resistant Pseudomonas putida DOT-T1E, three efflux pumps have been identified. The TtGABC and TtgGHI pumps extruded toluene, m-xylene, ethylbenzene, and styrene, whereas TtgDEF extruded only toluene and styrene (Rojas et al. 2001). In addition, mutants lacking the TtgABC pump displayed reduced resistance to nalidixic acid, chloramphenicol, and tetracycline, whereas a mutant lacking Ttg GHI was as sensitive as the parent strain. An efflux system has been characterized in a strain of Pseudomonas fluorescens cLP6a (Hearn et al. 2003). This system is selective for extrusion of phenanthrene, anthracene, and fluoranthene, whereas it is inactive for toluene and naphthalene. Gram-Positive Bacteria

Hydrocarbon tolerance has also been found in Gram-positive rhodococci. Tolerance to high concentrations of benzene has been demonstrated in a strain of Rhodococcus that is, in addition, tolerant of pHs in the range 2–10 (Paje et al. 1997). For Rhodococcus opacus, resistance to benzene, toluene,

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phenol, and chlorobenzene was accompanied by an increase in the synthesis of 10-methyl branched fatty acids at the putative expense of their unsaturated fatty acids (Tsitko et al. 1999). A number of factors may therefore be involved in the tolerance of bacteria to aromatic hydrocarbons, although greatest attention has hitherto been centered on Gram-negative organisms. Antibiotic Resistance Strains of Streptomyces isolated from a range of urban, agricultural, and forest soils isolated on a standard medium displayed resistance to antibiotics that were well established in clinical practice. Resistance was found among all classes of antibiotics, and a disturbing number of isolates were resistant to between 5 and 10 of them (D’Costa et al. 2006). Resistance to a range of antibiotics is of increasing concern in clinical practice since the genes are often carried on transmissible plasmids. There are different types of mechanism that confer resistance, including enzymatic covalent modification of the antibiotic, effective efflux systems, and induction of a cellular enzyme that is resistant to the antibiotic. Examples of these are used as illustration. Enzymatic Covalent Modification of Antibiotic—The Range of Reactions 1. Hydrolysis of penicillin mediated by penicillinase that has motivated attempts to alter its structure (e.g., in semisynthetic penicillins). 2. O-acetylation of chloramphenicol (Shaw and Leslie 1991) and zwittermicin A (Stohl et al. 1999). For chloramphenicol, the reaction is complex involving three reactions: (i) acetylation at C-3, (ii) nonenzymatic acetyl migration to C-1, followed by (iii) acetylation at C-3 to the 1,3-diacetyl compound. 3. O-phosphorylation of (a) chloramphenicol (Mosher et al. 1995) and spectinomycin that lead to resistance in Legionella pneumophila (Thompson et al. 1998) and (b) erythromycin at the 2′ position (Noguchi et al. 1995). 4. O-glucosylation of telithromycin at the 2′ position in Streptomyces sp. strain Ja#7 (D′Costa et al. 2006). 5. Reduction of 5-nitrofuran derivatives associated with mutation in aryl nitroreductases (Whiteway et al. 1998). 6. Monooxygenation of tetracycline by TetX (Yang et al. 2004). In contrast, resistance to isonicotinic acid hydrazide (INH) that is the frontline drug against Mycobacterium tuberculosis and the M. tuberculosis group—though not against M. leprae—illustrates the converse. INH is a prodrug that is converted into the active form, and resistance is achieved by mutation of the gene KatG that encodes a catalase-peroxidase (Bertrand et al. 2004). Details of the mechanism of activation are complex and two possibilities have been considered: 1. Activation by KatG to a range of products including isonicotinic acid and isonicotinaldehyde. This is supported by, for example, the mechanism of enzyme-catalyzed activation in peroxidases (Pierattelli et al. 2004), and the analogous activation of ethionamide by S-monooxygenation (DeBarber et al. 2000). Remarkably, the second-tier antituberculosis drug ethionamide is also a prodrug that is oxidized by an FAD-containing enzyme (EtaA) to the S-oxide and then to the amide (Vannelli et al. 2002). The initial reaction carried out by EtaA is a Baeyer–Villiger monooxygenation that functions typically with other substrates including 4-thiomethyltoluene and ketones (Fraaije et al. 2004). 2. The role of NO is supported by several lines of evidence: a. Its formation has been shown by electron paramagnetic spin resonance using 15N-labelled INH, and CPTIO, which is a specific scavenger of NO-provided protection against INH activity in liquid cultures of Mycobacterium tuberculosis (Timmins et al. 2004).

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b. Formation of nitrotyrosine by nitric oxide carriers during oxidation of INH (Van Zyl and van der Walt 1994). c. Incorporation into ciprofloxacin of an ester with a terminal –O–NO2 group that can release NO stimulates its activity against M. tuberculosis (Ciccone et al. 2003). An alternative and not necessarily conflicting view is the formation of a free radical at a tyrosine center that was trapped by adding NO to form 3-nitrotyrosine (Zhao et al. 2004). Efflux Systems An alternative to most of these mechanisms is the existence of efficient efflux systems, so that toxic concentrations of the drug are not achieved. There are three major families of proton-dependent multidrug efflux systems: (1) the major facilitator superfamily, (2) the small multidrug resistance family, and (3) the resistance/nodulation/cell division family (Paulsen et al. 1996). It should be emphasized that several of these systems are involved not with antibiotic efflux but with, for example, acriflavine, chlorhexidine, and crystal violet. An attempt is made only to outline a few salient features of the resistance/nodulation/cell division family that mediates antibiotic efflux, and these are given in Table 3.3 (Nikaido 1996). They consist of a transporter, a linker, and an outer membrane channel. Alteration of an Essential Reaction that is Normally Inhibited The macrolide erythromycin inhibits protein synthesis and resistance is induced by N6 -dimethylation of adenine within the 23S rRNA, which results in reduced affinity of ribosomes for antibiotics related to erythromcin (Skinner et al. 1983). Sulfonamides function by binding tightly to chromosomal dihydropteroate synthetase and resistance to sulfonamides is developed in the resistance plasmid through a form of the enzyme that is resistant to the effect of sulfonamides. Antifungal Agents This is not discussed in detail since mechanisms of resistance have been carefully reviewed (Ghannoum and Rice 1999). It was pointed out that resistance has not been associated with modification of the structure. For the 1,2,4-triazoles that have been widely used, their effect is due to inhibition of the synthesis of ergosterol that is the dominant component of fungal cell membranes. Resistance is generally associated with modification of the target enzymes, for example, the epoxidation of squalene (Terbinafine) or 14α-demethylase (Fluconazole). Resistance of Candida albicans to the azole antifungal agent fluconazole demonstrated, however, the simultaneous occurrence of several types of mechanism for resistance (Perea et al. 2001): a. Levels in expression of the genes encoding lanosterol 14α-demethylase (ERG11) both by overexpression that necessitates higher intracellular concentration of the drug and by point mutations that result in diminished affinity for the drug. b. The existence of multidrug efflux transporters including both MDR1 and CDR. TABLE 3.3 Summary of Resistance/Nodulation/–Cell Division Family Transporter

Linker

Outer Membrane Channel

Organism

Substrates

AcrB MexB MtrD

AcrA MexA MtrC

TolC OprM MtrE

E. coli P. aeruginosa Neisseria gonorrhoeae

TC, CA, FQ, ERY, NOV, RIF, β-LAC TC, CP, FQ, β-LAC TC, CP, β-LAC, RIF

Note: TC, tetracycline; CA, chloramphenicol; FQ, fluoroquinoline; ERY, erythromycin; NOV, novobiocin; RIF, rifampin; and β-LAC, β-lactam.

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Many yeasts are inhibited by 5-fluorocytosine and a block in the synthesis of 5-fluorouridylic acid by loss of cytosine deaminase or of uracil phosphoribosyltransferase is sufficient to cause resistance. Mutational loss of pyrimidine salvage enzymes has been frequently observed.

RESISTANCE TO METALS AND METALLOIDS There is concern over the toxicity of a number of metals and metalloids, and their oxyanions. As for antibiotics, the genes for resistance are often plasmid-borne. There are several mechanisms that may operate—reduction, methylation, efflux, and the synthesis of metal-binding metallothioneins. The following text illustrates aspects of these mechanisms. Reduction Reduction to less toxic forms, for example, Hg2+ to Hg0, CrO42− to Cr3+, U(VI) to U(IV), SeO32− to Se0, TeO32− to Te0, and AsO43− to AsO33−. Extrusion of the reduced form may also be involved in, for example, Se0, Te0, and AsO32− so that synthesis of the more toxic AsO32− is circumvented. Mercury Mercuric reductase is a key enzyme in detoxifying inorganic mercury (Hg2+) by reducing it to the nontoxic Hg0. The reductase in a range of rapid-growing mycobacteria from clinical sources increased from 20- to >100-fold after exposure to HgCl2 (Steingrube et al. 1991). The reductases from the plasmid-carrying Escherichia coli strain J53-1(R831) (Schottel 1978) and from Pseudomonas aeruginosa carrying the plasmid pVS1 (Fox and Walsh 1982) have been purified and both contained FAD. The enzyme from the latter was shown to contain two thiol groups that are released after reduction with NADPH, and was considered to be similar to lipoamide dehydrogenases. The degradation of phenylmercuric acetate to benzene, methylmercuric chloride to methane, and ethylmercuric chloride to ethane and Hg2+ is apparently carried out by different enzymes from the plasmid-carrying Escherichia coli strain K12 (R831) (Schottel 1978) and Pseudomonas sp. Resistance to organic mercury compounds has also been found in clinical isolates of nontuberculous, rapidly growing mycobacteria (Steingrube et al. 1991) and can present a challenge in the clinical environment. Chromium Although reduction of chromate CrVI to CrIII has been observed in a number of bacteria, these are not necessarily associated with chromate resistance. For example, reduction of chromate has been observed with cytochrome c3 in Desulfovibrio vulgaris (Lovley and Phillips 1994), soluble chromate reductase has been purified from Pseudomonas putida (Park et al. 2000), and a membrane-bound reductase has been purified from Enterobacter cloacae (Wang et al. 1990). The flavoprotein reductases from Pseudomonas putida (ChrR) and Escherichia coli (YieF) have been purified and can reduce Cr(VI) to Cr(III) (Ackerley et al. 2004). Whereas ChrR generated a semiquinone and reactive oxygen species, YieR yielded no semiquinone, and is apparently an obligate four-electron reductant. It could therefore present a suitable enzyme for bioremediation. Arsenic and Antimony A number of redox systems are involved in microbial reactions that confer resistance to inorganic arsenic (Silver and Phung 2005). The arsenate reductase gene ArsC both in Staphylococcus aureus (Ji and Silver 1992) and in Escherichia coli (Liu et al. 1995) is a determinant of arsenate resistance, and the arsenite that is produced is extruded from the cell by the ArsA–ArsB anion-translocating ATPase (Walmsley et al. 1999). In contrast, ArsB catalyzes arsenite extrusion coupled to electrochemical energy, and transport of Sb(III) into E. coli is catalyzed by the ArsB carrier protein (Meng et al. 2004). Another mechanism for detoxification of As(III) has been found in the legume symbiont Sinorhizobium meliloti. In the operon that is involved in As(III) detoxification, the arsb gene is replaced by aqpS that encodes an aquaglyceroporin. AqpS therefore confers resistance to arsenate

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by facilitated extrusion of arsenite produced by ArsC-catalyzed reduction of arsenate (Yang et al. 2005). In contrast, it has been proposed that in Escherichia coli the polyol transporter GlpF accumulates both arsenite and the analogous antimonite. Arsenite is also an intermediate in the fungal biomethylation of arsenic (Bentley and Chasteen 2002) and oxidation to the less toxic arsenate can be accomplished by heterotrophic bacteria including Alcaligenes faecalis. Exceptionally, arsenite can serve as electron donor for chemolithotrophic growth of an organism designated NT-26 (Santini et al. 2000), and both selenate and arsenate can be involved in dissimilation reactions as alternative electron acceptors. In a wider context, reduction of Sb(V) in the native drug to Sb(III) is required for the activity of preferred drugs for the treatment of leishmaniasis, and is supported by the following evidence. The sequence for the arsenate reductase ScAcr2p in Saccharomyces cerevisiae was used to clone the homolog LmACR2 in Leishmania major, and it was shown that the purified enzyme was able to reduce both As(V) and Sb(V). In addition, transfection of L. infantum with LmACR2 increased the sensitivity of amastigotes to the Sb(V) drug Pentostam (Zhou et al. 2004). Selenium and Tellurium Whereas selenium is a necessary trace element particularly for anaerobic bacteria, tellurium is not known to play an essential role. Both selenite (SeO32−) and tellurite (TeO32−) can be toxic, and in Escherichia coli, selenite is able to induce the genes for Mn and Fe superoxide dismutases that may be protective under aerobic conditions (Bébien et al. 2002). Resistance to selenite and tellurite can be developed in a range of bacteria. A number of controversial mechanisms have been proposed for the reduction of SeO32− to Se0, and the involvement of glutathione has been suggested on the plausible basis of a chemical analogy (Kessi and Hanselmann 2004). In resistant bacteria related to the genus Pseudoalteromonas (Rathgeber et al. 2002) and in Shewanella oneidensis (Klonowska et al. 2005), the selenite that has been reduced to Se0 can be extruded from the cells, whereas the Te0 that is produced from tellurite is located within the cells. As for the mechanism of reduction, several reasons have been adduced for resistance to selenite and tellurite. In Escherichia coli, tellurite resistance has been associated with the activity of nitrate reductase (Avazéri et al. 1997; Sabaty et al. 2001). The ars operon on the plasmid R773, which functions as an anion-translocating ATPase for arsenite, arsenate, and antimonite, provides moderate resistance also to tellurite (Turner et al. 1992). In the Gram-positive Geobacillus stearothermophilus, which is naturally resistant to tellurite, transfer of the iscS gene that encodes cysteine desulfurase into E. coli confers tellurite resistance (Tantaleán et al. 2003). Detoxification of tellurite has also been attributed to methyltransferase activity involving S-adenosymethionine (Liu et al. 2000) or thiopurine methyltransferase (Cournoyer et al. 1998), even when the expected methylated products were not observed (Liu et al. 2000). Selenate can be reduced to Se0 via selenite in Enterobacter cloacae SLD1a-1 (Losi and Frankenberger 1997), although it cannot be used as an electron acceptor for anaerobic growth (Ridley et al. 2006). Selenate can also serve as an electron acceptor, which is discussed in Part 2 of this chapter. Although in Rhodobacter sphaeroides (Van Fleet-Stalder et al. 2000) and in the metal-resistant Ralstonia metallidurans CH34 (Sarret et al. 2005), both reduction of selenite to Se0 and formation of organic selenium may take place, reduction appears to be the dominant reaction. Methylation Methylation of both metals and metalloids has been observed for both fungi and bacteria. These metabolites may, however, be toxic to higher biota as a result of their volatility. The Minamata syndrome represents the classic example of the toxicity of forms of methylated Hg to man, even though the formation of Hg(CH3)2 was probably the result of both biotic and abiotic reactions. Transmethylation is important not only in the biosynthesis of cellular components but also in detoxification. Although the resulting metabolites are less toxic to the cell, they are often more lipophilic and may have serious adverse effects on other biota. Toxic volatile arsenic compounds

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produced from inorganic. As were among the first examples, and methylmercury is a well-established toxicant both to wildlife and humans. Transmethylation is also involved in the degradation of methyl halides by the corrinoid pathway. Alkylated compounds were previously of industrial importance, for example, tetraethyl lead as an antiknock in gasoline engines and butyltin compounds as biocides. Biogeochemical cycles must take into account biogenic and anthropogenic inputs as well as biogenic and abiotic degradation and transformation. Monitoring has revealed the presence in the environment of a wide variety of methylated metals and metalloids. In the following summary, however, attention is directed only to results that conclusively demonstrate methylation under controlled laboratory conditions. Reference should be made to substantive reviews with extensive bibliographies dealing with methylation of metals and metalloids (Thayer 2002; Chasteen and Bentley 2002; Bentley and Chasteen 2003). A cardinal issue is the species of the metal or metalloid that is examined. Metals such as mercury or tin are methylated from cationic Hg2+ or Sn4+, whereas the metalloids are transformed from the oxyanions of As, Sb, Se, or Te. The classical Challenger mechanism that involves sequential reductions and methylations is well established, at least for fungal methylation of the oxyanions of As (Bentley and Chasteen 2002), and Se—and is assumed to be—for Te (Chasteen and Bentley 2003). Methylation may take place under aerobic conditions for fungi or anaerobic conditions for bacteria. Formation of methylated derivatives may occur in metals and metalloids belonging to groups 15 and 16, and a few of group 14 of the periodic table, and have been discussed in a critical review (Thayer 2002). Although bacteria can carry out several of the methylations, fungal methylation is probably most widely distributed, for example, of As and Sb (Tamaki and Frankenberger 1992; Bentley and Chasteen 2002), and Se and Te (Chasteen and Bentley 2003). Although the mechanism has not been established for all methylations, methyl transferases generally use either S-adenosylmethionine or methylcobalamin as methyl donors. In contrast, methylation of selenite in Escherichia coli is carried out by the activity of thiopurine methyltransferase (Ranjard et al. 2003). Methylation of mercury seems, however, to be carried out only by bacteria, although the mechanisms for methylation under anaerobic conditions have not been finally resolved (Bentley and Chasteen 2002). Sulfate-reducing bacteria are generally presumed to be the primary source of HgCH3 under anaerobic conditions (Choi et al. 1994; King et al. 2000), and generally use methylcobalamin as methyl donor. Methylation may also take place by reactions, which do not involve methylcobalamin (Ekstrom et al. 2003). It has been shown that Geobacter strain CLFeRB was able to carry out methylation of Hg2+ at a rate comparable to that of Desulfobulbus propionicus strain Ipr3 (Fleming et al. 2006). Under anaerobic conditions, methanogenic archaeons use pathways that probably involve methylcobalamin for the methylation of arsenate to dimethylarsine (McBride and Wolfe 1971); arsenate to mono-, di-, and trimethylarsines; SbCl3 to trimethylstibine; and Bi(NO3)3 to trimethylbismuth (Michalke et al. 2000). Conversely, among the organic arsenic compounds that were evaluated, only methylarsine appreciably inhibited acetoclastic methanogenesis (SierraAlvarez et al. 2004). Under aerobic conditions, S-adenosylmethionine is the methyl donor for methylation of methanethiol and methaneselenol (Drotar et al. 1987), and probably for the bacterial methylation of halogenated phenols and thiophenols (Neilson et al. 1988). It is also the probable methyl donor for fungal methylation of the oxyanions of As and Sb (Bentley and Chasteen 2002). Although the pathway has not been established, relatively high yields of trimethyltin from inorganic tin have been observed in yeast concomitant with the degradation of butyltin compounds (Errécalde et al. 1995). Exceptionally, methionine transferase may carry out the methylation of Hg in Neurospora crassa (Landner 1971) and thiopurine methyltransferase the methylation of inorganic Se in Escherichia coli (Ranjard et al. 2003). Although their source has not been identified, it has been shown that volatile compounds from landfills contain carbonyls of Mo and W in addition to the known hydrides and methylated derivatives of As, Se, Sn, Sb, Te, Hg, Pb, and Bi (Feldmann and Cullen 1997).

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Efflux Systems As for antibiotics, resistance to toxic cations may be mediated by the existence of efflux systems that have been discussed in reviews (Silver and Phung 1996; Mergeay et al. 2003; Nies 2003). A great deal of attention has been directed to Ralstonia metallidurans that has masqueraded under several names—Alcaligenes eutropha, Ralstonia eutropha, and Achromobacter xylosoxidans—and details of the mechanisms have been presented in a review (Mergeay et al. 2003). In this brief summary of efflux systems that alleviate heavy metal toxicity, the convenient system of classification used by Nies (2003) will be adopted. The range of metals in this section is restricted to resistance to the divalent cations of Zn, Cd, Co, and Ni, and the monovalent Cu and Ag, and occurs in a range of bacteria and a few yeasts. Mechanisms for resistance to mercury have already been discussed, and chromate resistance in Pseudomonas aeruginosa is achieved by an efflux system that is driven by the membrane potential (Alvarez et al. 1999). a. The resistance–nodulation–cell division (RND) family of proteins. In Ralstonia metallidurans the CzcA protein mediates resistance to Zn2+, Cd2+, Co2+, and Ni2+ in two megaplasmids, one associated (pMOL30) with resistance to Cd2+, Zn2+, and Co2+, the other (pMOL28) with increased resistance to Co2+ and Ni2+. In Escherichia coli, resistance to copper is mediated by a periplasmic protein (CusF). b. Cation-diffusion facilitators (CDF). The primary substrate for the protein is Zn2+, although other divalent cations including Ni2+ and Co2+, may also be effective. Transport is driven by gradients of several types—concentration, chemiosmotic ∆ψ or ∆pH, or potassium, and most studies have examined R. metallidurans in which the encoding genes are chromosomal, although species of Saccharomyces have also been included. c. P-type ATPases. These function both as uptake and as efflux mediators, and one family is involved in the specific efflux of monovalent Cu+ and Ag+, but not Cu2+, and has been examined in Gram-positive bacteria and the archaeon Archaeoglobus fulgidus. Another family that is involved in efflux of Zn2+/Cd2+/Pb2+ has been found in Gram-negative bacteria and yeast. Polyphosphates There is some evidence for the role of phosphate produced by hydrolysis of polyphosphates in metal tolerance and remediation. Polyphosphates are accumulated by many bacteria and serve as an energy source and for chelating metal cations. Their biosynthesis is mediated by a kinase ppk that catalyzes their formation from the terminal phosphate of ATP, while an exopolyphosphatase ppx brings about their hydrolysis to inorganic phosphate. It has been suggested that the extruded phosphate could bind metal cations and therefore lead to the development of resistance to the toxic cation. For example, the thermophilic archaeon Sulfolobus metallicus was highly resistant to Cu2+ during the concomitant decrease in the level of polyphosphate, increase in the level of exopolyphosphatase, and the efflux of phosphate (Remonsellez et al. 2006). Overexpression of the polyphosphate kinase gene in Pseudomonas aeruginosa induced accumulation of polyphosphate that could be mobilized under conditions of carbon starvation with the release of phosphate into the medium. Phosphate was utilized to remove UO22+ from solution that was initially precipitated on the cell walls (Renninger et al. 2004). Metallothionein Metallothioneins are proteins that are rich in cysteine residues, and the genes encoding them are found in eukaryotes and a few prokaryotes. Although they have been found in a wide range of biota, their function has not been finally resolved, and different roles have been assigned to them (Palmiter 1998). To demonstrate their function in metal resistance, it is necessary to show not only their presence, but also that the genes encoding them are amplified after exposure to the metal. Resistance has been explored extensively for copper that can produce reactive oxygen species with

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consequent cellular damage. At the same time, copper is necessary for the functioning of several cellular enzymes. Homeostasis is maintained by regulation. Whereas this can be maintained in prokaryotes by P1-type ATPases in combination with a regulatory protein, for example, in Pseudomonas putida (Adaikkalam and Swarup 2002), tolerance may also be mediated by metallothioneins. They have been found in a few cyanobacteria within the genus Synechococcus (Turner and Robinson 1992), and the structure in strain Tx-20 has been elucidated (Olafson et al. 1988). It has been shown in a culture of Synechococcus PCC 6301 adapted to Cd2+ that amplification of the smtAB genes took place, and that multiple copies of them were found after exposure to Cd2+ to which the cells are resistant (Gupta et al. 1992). Genes homologous to these have also been found in Cd-resistant strains of sulfate-reducing bacteria. Metal-translocating ATPases were, however, also found in two of the strains (Naz et al. 2005), so that several mechanisms of resistance could be involved. Substantial studies have been directed to yeasts. Several studies have examined the resistance to Cu2+ and Ag+ in the yeast Candida. In C. albicans, genes encoding a P1-type ATPase, and one encoding a metallothionein were cloned, and their expression studied in response to Cu2+. Transcription of only the former increased with increasing levels of Cu2+, so that this was putatively the primary determinant of resistance (Riggle and Kumamoto 2000). A similar study showed that extracellular Cu2+ induced the genes for a metallothionein (CaCUP1) and a P-type ATPase (CaCRP1). The former was responsible for the unusual resistance of C. albicans to copper, whereas the latter was essential for survival under conditions of low copper concentration (Weissman et al. 2000). In Candida glabrata, whereas glutathione-related (γEC)nF peptides are induced by Cd2+, metallothioneins encoded by the genes MT-1 and MT-2 are induced by Cu2+. In cells induced by increasing concentrations of Cu, chromosomal amplification of the MT-II took place, whereas the MT-I gene remained as a single copy and was shown to map on different chromosomes (Mehra et al. 1990). Although protection from copper toxicity in Saccharomyces cerevisiae involves a metallothionein encoded by CUP1, an additional element encoded by CRS5 that is regulated by both Cu2+ and by oxidative stress possesses a metallothionein-like amino acid sequence (Culotta et al. 1994). It is therefore clear that care should be exercised in assigning metallothioneins to a cardinal role in conferring resistance to metals. Chlorophenol Tolerance There are two putative mechanisms whereby bacteria become resistant to polychlorinated phenols. These are illustrated by the following examples: 1. Two strains of Sphingomonas sp. that could degrade pentachlorophenol maintained their levels of ATP even in the presence of high concentrations of pentachlorophenol. Analysis of the lipids using 31P NMR showed that this could be attributed to the increased levels of cardiolipin (Lohmeier-Vogel et al. 2001). 2. O-methylation has been shown in a range of halogenated phenols and thiophenols (Neilson et al. 1988), thiophenols (Drotar and Fall 1985), and methanethiol and methaneselenol (Drotar et al. 1987). Methylation has been putatively attributed to reaction with S-adenosylmethionine. Resistance to High Acidity Some bacteria are able to withstand high acidity in their surroundings, and for some pathogens this is critical for their survival. Some details of this are given in Chapter 7, Part 1, and it is sufficient here to merely note the existence of mechanisms whereby the decarboxylation of aminoacids serves to remove protons from the medium.

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Santini JM, LI Sly, RD Schnagl, JM Macy (2000) A new chemolithoautotrophic arsenite-oxidizing bacterium isolated from a gold mine: phylogenetic, physiological and preliminary biochemical studies. Appl Environ Microbiol 66: 92–97. Sarret G, L Avoscan, M Carrière, R Collins, N Geoffroy, F Carrot, J Covès, B Gouget (2005) Chemical forms of selenium in the metal-resistant bacterium Ralstonia metallireducens CH34 exposed to selenite and selenate. Appl Environ Microbiol 71: 231–2337. Schottel JL (1978) The mercuric and organomercurial detoxifying enzymes from a plasmid-bearing strain of Escherichia coli. J Biol Chem 253: 4341–4349. Shaw WV, AGW Leslie (1991) Chloramphenicol acetyltransferase. Annu Rev Biophys Chem 20: 363–386. Sierra-Alvarez R, I Cortinas, U Venal, JA Field (2004) Methanogenesis inhibited by arsenic compounds. Appl Environ Microbiol 70: 5688–5691. Sikkema J, JAM de Bont, B Poolman (1995) Mechanisms of membrane toxicity of hydrocarbons. Microbiol Rev 59: 201–222. Silver S, LT Phung (1996) Bacterial heavy metal resistance: new surprizes. Annu Rev Microbiol 50: 753–789. Silver S, LT Phung (2005) Genes and enzymes in bacterial oxidation and reduction of inorganic arsenic. Appl Environ Microbiol 71: 599–608. Skinner R, E Cundliffe, FJ Schmidt (1983) Site of action of a ribosomal RNA methylase responsible for resistance to erythromycin and other antibiotics. J Biol Chem 258: 12702–12706. Steingrube VA, RJ Wallace, LC Steele, Y Pang (1991) Mercuric reductase activity and evidence of broadspectrum mercury resistance among clinical isolates of rapidly growing mycobacteria. Antimicrob Agents Chemother 35: 819–823. Stohl EA, SF Brady, J Clarty, J Handelsman (1999) ZmaR, a novel and widespread antibiotic resistance that acetylates zwittermicin A. J Bacteriol 181: 5455–5460. Tamaki S, WT Frankenberger (1992) Environmental biochemistry of arsenic. Rev Environ Contam Toxicol 124: 79–110. Tantaleán JC, MA Araya. CP Saavedra, DE Fuementes, JM Pérez, IL Calderón, P Youderian, CC Vásquez (2003) The Geobacillus stearothermophilus V iscS gene, encoding cysteine desulfurase, confers resistance to potassium tellurite in Escherichia coli. J Bacteriol 185: 5831–5837. Thayer JS (2002) Biological methylation of less-studied elements. Appl Organometal Chem 16: 677–691. Thompson PR, DW Hughes, NP Cianciotto, GD Wright (1998) Spectinomycin kinase from Legionella pneumoniae. J Biol Chem 273: 14788–14795. Timmins GS, S Master, F Rusnak, V Deretic (2004) Nitric oxide generated from isoniazid activation by KatG: source of nitric oxide and activity against tunerculosis. Antimicrob Agents Chemother 48: 3006–3009. Tsitko IV, GM Zaitsev, AG Lobanok, MS Salkinoja-Salonen (1999) Effect of aromatic compounds on cellular fatty acid composition of Rhodococcus opacus. Appl Environ Microbiol 65: 853–855. Turner JS, NJ Robinson (1992) Cyanobacterial metallothioneins: biochemistry and molecular genetics. J Ind. Microbiol 14: 119–125. Turner RJ, Y Hou, JH Weiner, DE Taylor (1992) The arsenical APTase efflux pump mediates tellurite resistance. J Bacteriol 174: 3092–3094. Van Fleet-Stalder V, TG Chasteen, IJ Pickering, GN George, RC Prince (2000) Fate of selenate and selenite metabolized by Rhodobacter sphaeroides. Appl Environ Microbiol 66: 4849–4853. Van Zyl JM, BJ van der Walt (1994) Apparent hydroxyl radical generation without transition metal catalysts and tyrosine nitration during oxidation of the anti-tubercular drug, isonicotinic acid hydrazide. Biochem Pharmacol 48: 2033–2042. Vannelli TA, A Dykman, PR Ortiz de Montellano (2002) The antituberculosis drug ethionamide is activated by a flavin monooxygenase. J Biol Chem 277: 12824–12829. Walmsley AR, T Zhou, MI Boirges-Walmsley, BP Rosen (1999) The ATPase mechanism of ArsA, the catalytic subunit of the arsenite pump. J Biol Chem 274: 16153–16161. Wang P, T Mori, K Toda, H Ohtake (1990) Membrane-associated chromate reductase activity from Enterobacter cloacae. J Bacteriol 172: 1670–1672. Weber FJ, S Isken, JAM de Bont (1994) Cis/trans isomerization of fatty acids as a defence mechanism of Pseudomonas putida strains to toxic concentrations of toluene. Microbiology (UK) 140: 2013–2017. Weissman Z, I Berdicevsky, BZ Cavari, D Kornotzer (2000) The high copper tolerance of Candida albicans is mediated by a P-type ATPase. Proc Natl Acad Sci USA 97: 3520–3525 Whiteway J, P Koziarz, J Veall, N Sandhu, P Kumar, B Hoecher, IB Lambert (1998) Oxygen-insensitive nitroreductases: analysis of the roles of nfsA and nfsB in development of resistance to 5-nitrofuran derivatives in Escherichia coli. J Bacteriol 180: 5529–5539.

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PART 5 ENZYMES CONTAINING MANGANESE, IRON, NICKEL, COPPER, MOLYBDENUM, TUNGSTEN, AND VANADIUM This short section attempts to bring together the range of metalloenzymes that are encountered in biodegradation and biotransformation. Fe is the most common component of enzymes, and is followed in frequency by zinc and molybdenum, while some important enzymes contain nickel, copper, manganese, tungsten, or vanadium.

MANGANESE The role of manganese concentration has seldom been explicitly examined in the context of biodegradation. It is a component of the oxygen-evolving complex containing the Mn4Ca cluster in oxygenic phototrophs, and it is essential for the growth of the purple nonsulfur anaerobic phototrophs Rhodospirillum rubrum and Rhodopseudomonas capsulata during growth with N2—though not with glutamate (Yoch 1979). 1. A few extradiol dioxygenases contain Mn in place of Fe: a. 3,4-Dihydroxyphenylacetate-2,3-dioxygenase from Arthrobacter globiformis (Boldt et al. 1995) and from Bacillus brevis (Que et al. 1981) are Mn(II) enzymes, and are neither activated by Fe(II) nor rapidly inhibited by H2O2. b. The 2,3-dihydroxybiphenyl dioxygenase from Bacillus sp. strain JF8 that mediates fission of the biphenyl ring is unusual in being Mn(II)-dependent, and differs in structure from the analogous enzymes in Burkholderia sp. strain LB400 and Pseudomonas paucimobilis strain KF707 (Hatta et al. 2003). 2. The aerobic and anaerobic degradation of acetone is initiated by carboxylation to acetoacetate. The involvement of manganese has been examined in photoheterotrophically grown Rhodobacter capsulatus strain B10 and the presence of Mn verified from the X-band EPR spectrum (Boyd et al. 2004). 3. There are several peroxidases produced by fungi including manganese peroxidase. This plays an essential role in the metabolic capability of the white-rot fungus Phanaerochaete chrysosporium. Two groups of peroxidases are produced during secondary metabolism— lignin peroxidases and manganese-dependent peroxidases. Both are synthesized when only low levels of Mn(II) are present in the growth medium, whereas high concentrations of Mn result in the repression of the synthesis of the lignin peroxidases and an enhanced synthesis of manganese-dependent peroxidases (Bonnarme and Jeffries 1990; Brown et al. 1990). Experiments with a nitrogen-deregulated mutant have shown that nitrogen regulation of both these groups of peroxidases is independent of Mn(II) regulation (van der Woude 1993).

IRON Cells require Fe, which is a component of cytochromes and of enzymes that carry out degradation by oxygenation. Its concentration within the cell must, however, be controlled to avoid the

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production of deleterious hydroxyl radicals from molecular oxygen (Touati et al. 1995). Under Fe limitation, transport of Fe into Gram-negative bacteria is controlled by specific Fe(III) siderophores containing catechol, α-hydroxycarboxylate, and hydroxamate groups in which the Fe occurs as a hexadentate octahedral complex. These are synthesized within the cell and excreted into the medium. They are bound by specific outer membrane receptors and cross the cytosolic membrane potential using the energy-transducing system TonB-ExcB-ExbD. In Gram-positive bacteria, however, ABC permeases are used (Andrews et al. 2003; Wandersman and Delepelaire 2004). Some organisms including pathogenic bacteria and fungi are able to make use of heme as a source of Fe, and this is noted in Part 1 of this chapter.

NICKEL Nickel exists in the tunicate Trididemnum solidum as the nickel complex of a modified chlorin (Bible et al. 1988) and is a component of a number of enzymes. Urease is the classic example of a nickel-containing enzyme, and several enzymes contain both nickel and iron. Details of enzymes that contain nickel have been provided in a review (Mulrooney and Hausinger 2003), and only brief summaries are provided: 1. Urease catalyzes the two-stage hydrolysis of urea to ammonia and bicarbonate, and is involved in the pathogenesis of reactive arthritis following infection with Yersinia enterocolitica, in urinary tract infection, and in exacerbating the function of Helicobacter pylori by increasing the pH. The enzyme is found in several Enterobacteriaceae—including Escherichia coli, and species of Proteus and Providencia—in which it is induced by urea, although in Klebsiella aerogenes its synthesis is regulated by nitrogen limitation. The enzyme contains two nickel atoms coordinated with His134, 136, 246, and 272 and the carboxylate of Asp360 (Mobley et al. 1995). 2. Glyoxalase I is involved in the detoxification of methylglyoxal (Figure 3.33) that is produced during unregulated metabolism of carbohydrates. It is a widely distributed and is typically a Zn-containing enzyme. The enzyme from Escherichia coli is, however, activated by Ni (Clugston et al. 1998), and the basis of metal ion activation has been examined (He et al. 2000). 3. Superoxide dismutase is important for the detoxification of the superoxide radical (O2−) by reacting with protons to produce H2O2: 2O2− + 2H+ → O2 + H2O2. Although the enzyme generally contains Mn and Fe, or Cu and Zn, the enzyme from Streptomyces seoulensis contains Ni(III) (Wuerges et al. 2004). 4. Exceptionally, in Escherichia coli acireductone dioxygenase (enediol dioxygenase) carries out two enzymatic activities that are responsible for the salvage of methionine, but are encoded by the same gene. Whereas one enzyme is dependent on Fe and produces the ketoacid and formate (Figure 3.34a), the other that is nickel-dependent produces the carboxylic acid, formate, and CO (Figure 3.34b) (Dai et al. 1999). 5. Methyl coenzyme M reductase plays a key role in the production of methane in archaea. It catalyzes the reduction of methyl-coenzyme M with coenzyme B to produce methane and the heterodisulfide (Figure 3.35). The enzyme is an α2β2γ2 hexamer, embedded between two molecules of the nickel-porphinoid F430 and the reaction sequence has been delineated (Ermler et al. 1997). The heterodisulfide is reduced to the sulfides HS-CoB and HS-CoM by a reductase that has been characterized in Methanosarcina thermophila, and involves low-potential hemes, [Fe4S4] clusters, and a membrane-bound methanophenazine that contains an isoprenoid chain linked by an ether bond to phenazine (Murakami et al. 2001).

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GSH OH CH3−CO−CHO CH3−CO−CH SG

FIGURE 3.33

CH3

C

HO

C

R

R

C

O O H C

C

OH

O R

C

O H C O

OH

C

SG

C

O C

(a)

H

O SG

OH C

CH3

CO2 H

O

O R

CH(OH)

C

C

OH + H−CO2H

O

OH

O OH (b)

O R

FIGURE 3.34

CH3−CH(OH)−C

Degradation of methylglyoxal by glyoxylase I.

O

O

OH

O

O

CH(OH)

C C O OH

R

C

OH

+ CO + H−CO2H

(a,b) Alternative dioxygenations of an enediol.

O CH3

SO3

S

+ HS

N

H

H

H CO2

CH4 + S

SO3

H

S

N

CH3

P O

O H P O

CH4

S

CoM

FIGURE 3.35

CoB

S

H

CoM

S

S

H CO2 CH3

CoB

Biosynthesis of methane from S-methylcoenzyme M by methyl coenzyme reductase.

6. The acetyl-coenzyme A decarbonylase synthase complex contains five polypeptide subunits and in acetate-degrading methanotrophs, such as Methanosarcina barkeri and M. thermophila, catalyzes the formation of methane and CO2 from acetyl-CoA: CH3–CO–S–CoA + Co(I)-FeS-protein ⇔ CH3 –Co(III)-FeS-protein + CO + CoA The methyl group is subsequently transferred to a tetrahydropterin and coenzyme M. The β-subunit contains Ni and an Fe/S center, and an Ni2[4Fe–4S] arrangement at the active site has been proposed (Gencic and Grahame 2003). 7. Carbon monoxide dehydrogenase can participate in different reactions. Both aerobic and anaerobic organisms can oxidize CO using the reaction CO + H2O → CO2 + H2. Whereas the former uses an Mo-[2Fe–2S]-FAD enzyme, the enzyme from the latter contains Fe and Ni. Carbon monoxide dehydrogenase occurs in several anaerobic bacteria including the homoacetate-fermenting Clostridium thermoaceticum (Drake et al. 1980) and the acetate-utilizing methanogen Methanosarcina thermophila (Lu et al. 1994). It has been suggested that in Carboxydothermus hydrogenoformans there are two forms of the enzyme, both of which contain Ni, one involved in energy generation and one in biosynthetic reactions (Svetlitchnyi et al. 2001). Carboxydothermus hydrogenoformans contains several [4Fe–4S] metal clusters and the active site contains an asymmetrical [Ni–4Fe–5S] cluster in which three of the Fe atoms and the Ni atom are coordinated with cysteine (Dobbek

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(1)

(2)

(3)

Environmental Degradation and Transformation of Organic Chemicals

F + P

E

E

FIGURE 3.36

cys

cys

C

S

S

O NH 2

C

C

AMP

F

O E

NH 2

N

E

O

cys

cys

C

S

O E

NH 2

C

O P

E

NH

SH + Fe

C

cys

S

cys

C

S

O NH 2

C

N

N

Synthesis of Fe–C≡N ligand.

et al. 2001). In contrast, an [Ni–3Fe–4S] cluster was found in Rhodospirillum rubrum (Drennan et al. 2001). 8. Hydrogenase catalyzes the reaction H2 ⇔ 2H+ + 2e and is found in a diverse range of organisms: (a) anaerobes including species of Desulfovibrio, (b) the aerobe Rhizobium leguminosarum, (c) the facultative anaerobe Escherichia coli, and (d) hyperthermophilic archaea including Thermococcus kodakaraensis and Pyrococcus furiosus. There are three groups of hydrogenases distinguished by the presence of (i) Ni–Fe, (ii) Fe–Fe with a dinuclear iron center in anaerobic organisms, or (iii) unrelated with an Fe–S clusterfree enzyme. Both the Ni–Fe and Fe–Fe hydrogenases have CO and CN ligands at the active site. Exceptionally, in Desulfovibrio baculatus, the active site contained [Ni–Fe–Se] (Wang et al. 1992). Nickel is required for the anaerobic growth of Escherichia coli, and there are four [Ni–Fe] hydrogenases whose genes occur in operons (references in Blokesch et al. 2004). The hydrogenases fulfill a range of functions. Two of the operons hya and hyb are induced under anaerobic conditions and allow the cells to utilize hydrogen as a source of energy, while hyc is part of the formate hydrogen lyase that converts formate into CO2 and hydrogen. In addition, synthesis of the active nickel enzyme requires the cooperation from several other proteins that bind zinc at a site apparently different from the Ni-binding site (Blokesch et al 2004). In Ni–Fe hydrogenases, the Fe is coordinated with one CO and two CN ligands, and synthesis of the latter involves two proteins HypF and HypE that carry out the consecutive reactions (Figure 3.36) (Reissmann et al. 2003). The analogous active site in the Fe–Fe hydrogenase of the hyperthermophilic Thermotoga maritima contains CO and CN ligands at both Fe atoms that are linked by an additional CO and sulfur ligands, one of which contains the [4Fe–4S] center. The maturation protein HydF is able to bind GTP and catalyze GTP hydrolysis, and contains an [4Fe–4S] cluster with an unusual EPR signal, which suggests that only three cysteines are coordinated (Brazzolotto et al. 2006). In hyperthermophilic archaea, hydrogenases are also involved in the reduction of S0 to H2S (Laska et al. 2003; Ma et al. 2000). It has been suggested that the subunits of hydrogenase and sulfur reductase in the acidophilic Acidianus ambivalens are similar to their homologs from Wolinella succinogenes (Laska et al. 2003).

COPPER Copper is a component of several enzymes that have quite different catalytic properties: 1. Methane monooxygenase may exist in either soluble (sMMO) or particulate (pMMO) forms. These display different substrate ranges and different rates of transformation rates, and most methanotrophs express only the latter form of the enzyme (Hanson and Hanson 1996). The particulate form of methane monooxygenase contains copper, or both copper

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2.

3.

4.

5. 6.

7.

185

and iron, and the concentration of copper determines the catalytic activity of the enzyme (Sontoh and Semrau 1998). Although various dioxygenases generally contain Fe(II) or Fe(III), copper replaces Fe in the dioxygenase from Aspergillus japonicus that brings about fission of the C-ring of quercitin with the formation of carbon monoxide (Steiner et al. 2002). The carbon monoxide dehydrogenase of the aerobe Oligotropha carboxidovorans contains both Cu and Mo in the form of a cluster in which the Mo is bound to the thiol groups of molybdopterin cytosine nucleotide, and the Cu to cysteine residue in the form of a Cu–S–Mo(=O)OH cluster (Dobbek et al. 2002). Glyoxal oxidase that is required to produce the H2O2 required for the activity of lignin peroxidase carries out the reaction RCHO + O2 + H2O → RCO2H + H2O2, and requires Cu for full activity. It is a radical-copper oxidase closely resembling galactose oxidase (Whittaker et al. 1996). Copper is a component of the oxalate decarboxylase in Bacillus subtilis that converts oxalate into formate and CO2 (Tanner et al. 2001). The deamination of primary amines such as phenylethylamine by Escherichia coli (Cooper et al. 1992) and Klebsiella oxytoca (Hacisalihoglu et al. 1997) is carried out by an oxidase. This contains copper and topaquinone (TPQ), which is produced from tyrosine by dioxygenation. TPQ is reduced to an aminoquinol that in the form of a Cu(I) radical reacts with O2 to form H2O2, Cu(II), and the imine. The mechanism has been elucidated (Wilmot et al. 1999), and involves formation of a Schiff base followed by hydrolysis in reactions that are formally analogous to those involved in pyridoxal-mediated transamination. Superoxide dismutases may contain a range of metals: Mn, Fe, or both Cu and Zn, and representatives of all these are found in prokaryotes. The nickel enzyme is noted later.

MOLYBDENUM Molybdopterin Oxidoreductases The molybdenum cofactor (molybdopterin) consists of a dihydropteridine fused to a dihydropyran ring containing a −CH2OPO32− group and an enedithiol that is coordinated to the molybdenum. In prokaryotes, the cofactor is formed by coupling to a nucleoside phosphate (generally cytidine or guanidine) forming a pyrophosphate (Llamas et al. 2004) (Figure 3.37). The biosynthesis of molybdopterin is catalyzed by at least six gene products, and is initiated from guanosine triphosphate by guanosine hydrolase I that provides the pteridine nucleus, while the C5 of ribose and the purine C8 provide the dihydropyran ring. Molybdopterin is a component of four enzyme families all of which contain Mo(VI): the xanthine oxidase and the sulfite oxidase families with one molybdopterin and the DMSO family with two molybdopterins. There are a number of tungsten-containing enzymes with structures analogous

O H

NH 2

H

S Mo

N

N N

N H

NH 2

S

O

N O

O P O

O

O P

O

O

O

N

O OH OH

FIGURE 3.37

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Structure of molybdopterin conjugated with cytidine diphosphate.

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to molybdopterin in which W(VI) replaces Mo(VI) (Hille 2002). Molybdopterin is a structural part of the important oxidoreductases that are involved in the aerobic degradation of azaarenes. Using a bioinformatics screen, a molybdenum-containing oxidoreductase has also been found in the majority of Gram-negative bacteria (Loschi et al. 2004). It contains one molybdopterin, and the soluble subunit designated YedY functions as a reductase for TMAO and some sulfoxides. These oxidoreductases are widely used for the introduction of the oxygen atom from H2O into heteroarenes, especially azaarenes including pyridine, quinoline, pyrimidine, and purine, and they generally contain molybdenum. Heteroarenes The degradation of the five-membered heteroarenes furan-2-carboxylate and pyrrole-2-carboxylate is initiated by the corresponding hydroxylases (Hormann and Andreesen 1991; Koenig and Andreesen 1990, 1991). Pyridine

The degradation of several pyridine carboxylates is initiated by hydroxylation. In Arthrobacter picolinophilus, pyridine-2-carboxylate is hydroxylated to 3,6-dihydroxypyridine-2-carboxylate (as for 2-hydroxypyridine) and this is carried out by a particulate enzyme that introduces oxygen from H2O (Tate and Ensign 1974), has a molecular mass of 130 kDa, and contains Mo (Siegmund et al. 1990). Pyridine-4-carboxylate is hydroxylated by Mycobacterium sp. strain INA1 to the 2,6-dihydroxy-4-carboxylate. Two different hydroxylation enzymes were involved and were apparently Mo-dependent (Kretzer and Andreesen 1991). The degradation of nicotine by Arthrobacter nicotinovorans involves two dehydrogenations by enzymes containing Mo, Fe–S, and FAD (Baitsch et al. 2001). Quinoline

Experiments with H218O using Pseudomonas putida strain 86 (Bauder et al. 1990) showed that the oxygen incorporated into quinol-2-one originates from water. The oxidoreductases have been purified from a number of organisms that degrade quinoline including P. putida (Bauder et al. 1990), Rhodococcus sp. strain B1 (Peschke and Lingens 1991), Comamonas testosteroni (Schach et al. 1995), and Agrobacterium sp. strain 1B that degrades quinoline-4-carboxylate (Bauer and Lingens 1992). They have a molecular mass of 300–360 kDa, and contain per molecule, eight atoms of Fe, eight atoms of acid-labile S, two atoms of Mo, and two molecules of FAD. The organic component of the pterin molybdenum cofactor is generally molybdopterin cytosine dinucleotide (Hetterich et al. 1991; Schach et al. 1995). The metabolism of 2-methylquinoline in Arthrobacter sp. strain Rü 61a is comparable, with introduction of oxygen at C4 (Hund et al. 1990). The enzymes (oxidoreductases) that introduce oxygen into the azaarene rings in Rhodococcus sp. strain B1, Arthrobacter sp. strain Rü61a, and Pseudomonas putida strain 86 are virtually identical and, like those already noted have molecular masses of 300–320 kDa and contain Mo, Fe, FAD, and acid-labile sulfur (De Beyer and Lingens 1993). The enzymes from Comamonas testosteroni for hydroxylation of quinoline to quinol-2-one (quinoline 2-oxidoreductase) and the dioxygenase responsible for the introduction of oxygen into the benzenoid ring (2-oxo-1,2-dihydroquinoline 5,6-dioxygenase) have been described (Schach et al. 1995). Isoquinoline

The degradation of isoquinoline by Alcaligenes faecalis strain Pa and Pseudomonas diminuta strain 7 (Röger et al. 1990, 1995) is mediated by an oxidoreductase that produces 1,2-dihydroisoquinoline-1-one, followed by ring fission with the production of o-phthalate and oxidation to 3,4-dihydroxybenzoate (Figure 3.38). The oxidoreductase is purified and like most typical azarene oxidoreductases contains, per mole, 0.85 g atoms of Mo, 3.9 g atoms of Fe, and acid-labile S (Lehmann et al. 1994).

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187 NH 2

CO 2 H CO 2 H

N

NH

CO 2 H

CO 2 H

O

OH OH

FIGURE 3.38 Degradation of isoquinoline initiated by an oxidoreductase.

Reductases a. The dehydroxylase in Thauera aromatica that brings about dehydroxylation of 4-hydroxybenzoyl-CoA to benzoyl-CoA is a molybdenum–flavin–iron–sulfur enzyme (Breese and Fuchs 1998). b. The DMSO reductase family contains molybdenum generally in the form of bis (molybdopterin guanine nucleotide) (references in McEwan et al. 2002). In addition to DMSO reductase, they include TMAO reductase, (per)chlorate reductase, selenate reductase, and arsenate reductase that are able to couple the reduction to the generation of energy (dissimilatory reduction). They catalyze the following reactions that have been noted in Part 2 of this chapter: Me2SO → Me2S (DMSO reductase); Me3NO → Me3N (TMAO reductase); HCO2H → CO2 + H2 (formate dehydrogenase); ClO4− → ClO3− → ClO− ([per]chlorate reductase); SeO42− → SeO32− (selenate reductase); AsO2− → AsO43− (arsenite oxidase). The chlorate reductase has been characterized in strain GR-1 where it was found in the periplasm, is oxygen-sensitive, and contains molybdenum, and both [3Fe–4S] and [4Fe–4S] clusters (Kengen et al. 1999). The arsenate reductase from Chrysiogenes arsenatis contains Mo, Fe, and acid-labile S (Krafft and Macy 1998), and the reductase from Thauera selenatis that is specific for selenate, is located in the periplasmic space, and contains Mo, Fe, acid-labile S, and cytochrome b (Schroeder et al. 1997). In contrast, the membrane-bound selenate reductase from Enterobacter cloacae SLD1a-1 that cannot function as an electron acceptor under anaerobic conditions contains Mo and Fe and is distinct from nitrate reductase (Ridley et al. 2006). In the anaerobic photoheterotroph Rhodobacter (Rhodopseudomonas) spheroides, a molybdenum enzyme in which the metal is coordinated by two equivalents of a pyranopterin cofactor is involved in sequential oxidations and reductions involving Mo(VI), Mo(V), and Mo(IV) (Cobb et al. 2005).

TUNGSTEN Tungsten-containing enzymes have been found to mediate a variety of reactions both in aerobic and anaerobic bacteria, and their structure may plausibly be assumed to be analogous to the molybdopterins: 1. Acetylene hydratase from the anaerobe Pelobacter acetylenicus is a tungsten– iron–sulfur enzyme that resembles molybdopterin in which W replaces Mo (Meckenstock

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2.

3.

4.

5.

Environmental Degradation and Transformation of Organic Chemicals

et al. 1999), and catalyzes the addition of the elements of water to acetylene HC≡CH → CH2=CHOH≡CH3CHO. Oxidoreductases: A number of these have been characterized and include the following: a. Aldehyde oxidoreductase from Desulfovibrio gigas (Hensgens et al. 1995). b. An aldehyde dehydrogenase from Desulfovibrio simplex is stimulated by tungsten and oxidizes aliphatic and aromatic aldehydes using flavins (Zellner and Jargon 1997). c. Aldehyde oxidoreductase from Clostridium thermoaceticum (Strobl et al. 1992). Several tungsten-containing oxidoreductases have attracted attention in the biochemistry of the extreme thermophile Pyrococcus furiosus: (a) aldehyde ferredoxin oxidoreductase (Mukund and Adams 1991), (b) glyceraldehyde ferredoxin oxidoreductase (Mukund and Adams 1995), (c) formaldehyde ferredoxin oxidoreductase (Mukund and Adams 1993; Roy et al. 1999), (d) an aldehyde oxidoreductase with a broad substrate specificity that probably involves ferredoxin (Bevers et al. 2005), and (e) an enzyme that may be involved in the reduction of S0 (Roy and Adams 2002). Carboxylic acid reductase: Anaerobic reduction of aryl and alkyl carboxylates to the corresponding aldehydes has been demonstrated in a few tungsten-containing reductases and dehydrogenases. In Clostridium formicoaceticum the enzyme catalyzes the reduction of propionate, butyrate, and benzoate to the corresponding aldehydes RCO2H → RCHO. The enzyme contains, per dimer, 11 iron, 16 acid-labile sulfur, and 1.4 tungsten + molybdopterin (White et al. 1991). In addition, an Mo oxidoreductase has been described from Cl. formicoaceticum (White et al. 1993). Formate dehydrogenase from Moorella (Clostridium) thermoaceticum catalyzes the reduction of CO2 to formate and contains W, Se, Fe, and S (Yamamoto et al. 1983), and the FDH-II enzyme from Eubacterium acidamophilum that uses formate as electron donor during aminoacid fermentations contains W, Se, and Fe (Graentzdoerffer et al. 2003). The formate dehydrogenase involved in the oxidation of formate to CO2 in E. coli is, however, an Mo, Se, (Fe 4S4) enzyme (Boyington et al. 1997). Hydrogenase: Thermotoga maritima grows at the expense of carbohydrates producing lactate, acetate, CO2, and H2. Although production of the hydrogenase and its in vitro activity is greatly stimulated by the presence of tungstate, the purified enzyme contains only Fe and acid-labile S, and tungsten is absent (Juszczak et al. 1991). This is also true for pyruvate: ferredoxin oxidoreductase from the same organism.

VANADIUM 1. Three enzymes in Azotobacter vinelandii (Rüttimann-Johnson et al. 2003) and Rhodopseudomonas palustris (Oda et al. 2005) are capable of reducing dinitrogen: the nifencoded enzyme containing molybdenum and iron, the vnf that encodes a vanadium and iron enzyme, and the anf that is an iron-only enzyme. 2. Some haloperoxidases contain vanadium and a review of vanadium peroxidases has been given (Butler 1998). The structure of the vanadium enzyme in the terrestrial fungus Curvularia inaequalis has been determined by x-ray analysis (Messerschmidt et al. 1997), and the apochloroperoxidase possesses, in addition, phosphatase activity that can be rationalized on the basis of the isomorphism of phosphate and vanadate (Renirie et al. 2000).

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Meckenstock RU, R Krieger, S Ensign, PMH Kroneck, B Schink (1999) Acetylene hydratase of Pelobacter acetylenicus. Molecular and spectroscopic properties of the tungsten iron-sulfur enzyme. Eur J Biochem 264: 176–182. Messerschmidt A, L Prade, R Wever (1997) Implications for the catalytic mechanism of the vanadiumcontaining enzyme chloroperoxodase from the fungus Curvularia inaequalis by X-ray structures of the native and peroxide form. Biol Chem 378: 309–315. Mobley HLT, MD Island, RP Hausinger (1995) Molecular biology of microbial ureases. Microbiol Rev 59: 451–480. Mukund S, MWW Adams (1991) The novel tungsten-iron-sulfur protein of the hyperthermophilic archaebacterium, Pyrococcus furiosus, is an aldehyde ferredoxin oxidoreductase J Biol Chem 266: 14208–14216. Mukund S, MWW Adams (1993) Characterization of a novel tungsten-containing formaldehyde oxidoreductase from the extremely thermophilic archaeon, Thermococcus litoralis. J Biol Chem 268: 13592–13600. Mukund S, MWW Adams (1995) Glyceraldehyde-3-phosphate ferredoxin oxidoreductase, a novel tungstencontaining enzyme with a potential glycolytic role in the hyperthermophilic archaeon, Pyrococcus furiosus. J Biol Chem 270: 8389–8392. Mulrooney SB, RP Hausinger (2003) Nickel uptake and utilization by microorganisms. FEMS Microbiol Rev 27: 239–261. Murakami E, U Deppenmeier, SW Ragsdale (2001) Characterization of the intramolecular electron transfer pathway from 2-hydroxyphenazine to the heterodisulfide reductase from Methanosarcina thermophila. J Biol Chem 276: 2432–2439. Oda Y, SK Samanta, FE Rey, L Wu, X Liu, T Yan, J Zhou, CS Harwood (2005) Functional genomic analysis of three nitrogenase isoenzymes in the photosynthetic bacterium Rhodopseudomonas palustris. J Bacteriol 187: 7784–7794. Peschke B, F Lingens (1991) Microbial metabolism of quinoline and related compounds XII. Isolation and characterization of the quinoline oxidoreductase from Rhodococcus spec. B1 compared with the quinoline oxidoreductase from Pseudomonas putida 86. Biol Chem Hoppe-Seyler 370: 1081–1088. Que L, J Widom, RL Crawford (1981) 3,4-dihydroxyphenylacetate 2,3-dioxygenase : a manganese(II) dioxygenase from Bacillus brevis. J Biol Chem 256: 10941–10944. Reissmann S, E Hochleitner, H Wang, A Paschos, F Lottspeich, RS Glass, A Böck (2003) Taming of a poison: biosynthesis of the NiFe-hydrogenase cyanide ligands. Science 299: 1067–1070. Renirie R, W Hemrika, R Wever (2000) Peroxidase and phosphatase activity of active-site mutants of vanadium chloroperoxidase from the fungus Curvularia inaequalis. J Biol Chem 275: 11650–11657. Ridley H, CA Watts, DJ Richardson, CS Butler (2006) Resolution of distinct membrane-bound enzymes from Enterobacter cloacae SKLD1a-1 that are responsible for selective reduction of nitrate and selenate anions. Appl Environ Microbiol 72: 5173–5180. Röger P, A Erben, F Lingens (1990) Microbial metabolism of quinoline and related compounds IV. Degradation of isoquinoline by Alcaligenes faecalis Pa and Pseudomonas diminuta 7. Biol Chem Hoppe-Seyler 370: 1183–1189. Röger P, G Bär, F Lingens (1995) Two novel metabolites in the degradation pathway of isoquinoline by Pseudomonas diminuta 7. FEMS Microbiol Lett 129: 281–286. Roy R, MWW Adams (2002) Characterization of a fourth tungsten-containing enzyme from the hyperthermophilic Pyrococcus furiosus. J Bacteriol 184: 6952–6956. Roy R, S Mukund, GJ Schut, DM Dunn, R Weiss, MWW Adams (1999) Purification and molecular characterization of the tungsten-containing formaldehyde ferredoxin oxidoreductase from the hyperthermophilic archaeon Pyrococcus furiosus: the third of a putative five-member tungstoenzyme family. J Bacteriol. 181: 1171–1180. Rüttimann-Johnson C, LM Rubio, DR Dean, PW Ludden (2003) VnfY is required for full activity of the vanadium-containing dinitrogenase in Azotobacter vinelandii. J Bacteriol 185: 2383–2386. Schach S, B Tshisuaka, S Fetzner F Lingens (1995) Quinoline 2-oxidoreductase and 2-oxo-1,2-dihydroquinoline 5,6-dioxygenase from Comamonas testosteroni 63. The first two enzymes in quinoline and 3-methylquinoline degradation. Eur J Biochem 232: 536–544. Schroeder I, S Rech, T Krafft, JM Macey (1997) Purification and characterization of the selenate reductase from Thauera selenatis. J Biol Chem 272: 23765–23768. Siegmund I, Koenig K, Andreesen JR (1990) Molybdenum involvement in aerobic degradation of picolinic acid by Arthrobacter picolinophilus FEMS Microbiol Lett 67: 281–284.

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4

Determinants and Interactions

This section addresses the conditions under which the degradation of organic compounds is carried out, the factors that determine their success, and attempts to provide a background for experimental aspects, which are discussed in Chapters 5 and 6, illustrations of the reactions are given in Chapters 7 through 11.

SINGLE SUBSTRATES: SEVERAL ORGANISMS Cultures of a single microorganism may occur naturally only in circumstances where extreme selection pressure operates. Generally, however, many different organisms with diverse metabolic potential will exist side by side. Therefore, metabolic interactions are probably the rule rather than the exception in natural ecosystems and biological treatment systems. Both the nature and the tightness of the association may vary widely, and the degradation of a single compound may necessitate the cooperation of two (or more) organisms. Some well-defined interactions and the different mechanisms underlying the cooperation are illustrated by the following examples: 1. One of the organisms fulfills the need for a growth requirement by the other, for example, vitamin requirements of one organism that is provided by the other. Examples are provided by biotin in cocultures of Methylocystis sp. and Xanthobacter sp. (Lidstrom-O’Connor et al. 1983), and thiamin in cocultures of Pseudomonas aeruginosa and an undefined Pseudomonas sp. that degraded the phosphonate herbicide glyphosate (Moore et al. 1983). 2. One organism may be able to carry out only a single step in biodegradation. Many examples among aerobic organisms have been provided (Reanney et al. 1983; Slater and Lovatt 1984) so that this is probably a widespread situation. The following three examples are given as illustration: a. The degradation of parathion was carried out by a mixed culture of Pseudomonas stutzeri and P. aeruginosa (Daughton and Hsieh 1977) in which the 4-nitrophenol initially formed by the former is metabolized by the latter (Figure 4.1). b. The degradation of 4-chloroacetophenone was accomplished by a mixed culture of an Arthrobacter sp. and a Micrococcus sp. The first organism was able to carry out all the degradative steps except the conversion to 4-chlorocatechol of the intermediate 4-chlorophenol that is toxic to this organism (Havel and Reineke 1993). c. A consortium of a Flavobacterium sp. and a Pseudomonas sp. could carry out the aerobic degradation of polyethylene glycol. The latter was required for the degradation of the glycollate produced by the former (Kawai and Yamanaka 1986). HO–(CH2–CH2O)n –CH2–CH2–O–CH2–CH2OH → HO–(CH2–CH2O)n –CH2–CH2–O –CH2–CO2H → HO–(CH2–CH2O)n.CH2–CH2–OH + HO–CH2–CO2H

3. Two organisms are required to maintain the redox balance. Among anaerobic bacteria, hydrogen transfer is important since the redox balance must be maintained, and the hydrogen concentration in mixed cultures may be critical. Interspecies hydrogen transfer has been demonstrated especially among populations of rumen bacteria containing methanogens where the concentration of hydrogen must be limited for effective functioning of the consortia. Illustrative examples have been summarized (Wolin 1982), and some additional 193

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EtO P EtO

OH

S EtO EtO

P

O

NO 2 OH

OH OH

NO2

NO2

FIGURE 4.1 Degradation of parathion by a mixed culture of two pseudomonads.

comments on degradative reactions that are dependent on hydrogen transfer mediated by one of the cooperating organisms are added here. Stable metabolic associations generally between pairs of anaerobic bacteria have been termed syntrophs, and these are effective in degrading a number of aliphatic carboxylic acids or benzoate under anaerobic conditions. These reactions have been discussed in reviews (Schink 1991, 1997; Lowe et al. 1993) that provide lucid accounts of the role of syntrophs in the degradation of complex organic matter. Two examples are given here to illustrate the experimental intricacy of the problems besetting the study of syntrophic metabolism under anaerobic conditions: 1. Oxidation under anaerobic conditions of long-chain aliphatic carboxylic acids was established in syntrophic cultures of Clostridium bryantii, Desulfovibrio sp. (Stieb and Schink 1985); Syntrophomonas sapovorans, Methanospirillum hungatei (Roy et al. 1986); and Syntrophomonas wolfei in coculture with H2-utilizing anaerobic bacteria (McInerney et al. 1981). The role of the second syntroph was to metabolize the reducing equivalents produced by oxidation of the carboxylic acids. S. wolfei was subsequently, however, adapted to grow with crotonate in pure culture (Beaty and McInerney 1987), and this procedure was also used for Cl. bryantii. 16S rRNA sequence analysis was then used to show the close relationship of these two organisms and to assign them to a new genus Syntrophospora (Zhao et al. 1990). However, anaerobic oxidation of carboxylic acids with chain lengths of up to 10 has been demonstrated in pure cultures of species of Desulfonema (Widdel et al. 1983), and aliphatic hydrocarbons may be completely oxidized to CO2 by a sulfate-reducing bacterium (Aeckersberg et al. 1991, 1998). 2. There has been considerable interest in the anaerobic degradation of propionate that is a fermentation product of many complex substrates, and syntrophic associations of acetogenic and methanogenic bacteria have been obtained. During the metabolism of propionate in a syntrophic culture (Houwen et al. 1991), the methanogens serve to remove hydrogen produced during the oxidation of propionate to acetate. The growth of syntrophic propionate-oxidizing bacteria in the absence of methanogens has, however, been accomplished using fumarate as the sole substrate (Plugge et al. 1993). Fumarate played a central role in the metabolism of this organism since it is produced from propionate via methylmalonate and succinate, and is metabolized by the acetyl-CoA cleavage pathway via malate, oxalacetate, and pyruvate. Pure cultures of organisms that can oxidize propionate either in the presence of a methanogen or using sulfate as electron acceptor have been obtained. These include both Syntrophobacter wolinii and Syntrophobacter pfenigii (Wallrabenstein et al. 1995). The interaction of two organisms, therefore, is clearly not obligatory for the ability to degrade these carboxylic acids under anaerobic conditions.

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COMETABOLISM AND RELATED PHENOMENA In natural ecosystems, it is indeed seldom that either a pure culture or a single substrate exists. In general, several substrates will be present, and these will include compounds of widely varying susceptibility to microbial degradation. The phenomenon where degradation occurs in the presence of two substrates has been termed “cooxidation” or, less specifically, “cometabolism” or “concurrent metabolism.” Unfortunately, the term cometabolism (Horvath 1972) has been used in different and conflicting ways. Since the prefix “co” implies “together,” it should not therefore be used when only a single substrate is present. The term “biotransformation” is unambiguous, and seems appropriate and entirely adequate. Detailed discussions have been presented (Dalton and Stirling 1982), and some of the conflicting aspects have been briefly summarized (Neilson et al. 1985). Whereas probably most aerobic bacteria that have been used in experiments on biodegradation were obtained by elective growth using the chosen substrate as sole source of both carbon and energy, anaerobic bacteria are often more fastidious in their nutritional demands. The addition of nutrient supplements, such as yeast extract or ruminant fluid, may be needed to stimulate or maintain growth. Although the metabolic conclusions from such laboratory experiments are unambiguous, care should be exercised in uncritically extrapolating the results to natural ecosystems that are unlikely to provide such a nutritious environment. In such experiments, unequivocal results may often advantageously be obtained by using suspensions of washed cells. Cometabolism merits, however, careful analysis since important metabolic principles underlie most of the experiments, even though confusion may have arisen as a result of ambiguous terminology. An attempt is therefore made to ignore semantic implications and to adopt a broad perspective in discussing this environmentally important issue. A pragmatic point of view has been adopted, and the following examples attempt to illustrate the kinds of experiments, which have been carried out under various conditions.

INDUCTION OF CATABOLIC ENZYMES PRE-EXPOSURE TO AN ANALOG SUBSTRATE Organisms may be obtained after elective enrichment with a given substrate but are subsequently shown to be unable to use the substrate for growth, although they are able to accomplish its incomplete metabolism. These situations should rather be termed biotransformation. A typical example is the partial oxidation of 2,3,6-trichlorobenzoate to 3,5-dichlorocatechol (Figure 4.2) (Horvath 1971). The catabolic enzymes may, however, be induced by pre-exposure to an analog substrate. Cells may be grown with this before exposure to the xenobiotic. Although the xenobiotic is extensively degraded, it cannot be used alone to support the growth of cells. When both substrates are simultaneously present, the term cometabolism is appropriate, but this term is quite unjustifiable for the situation in which cells grown with a given inducing substrate are then used for studying the metabolism of a single xenobiotic. The following examples are given as illustration: a. Several strains of bacteria grown with nonchlorinated substrates such as phenol, naphth-2-ol, or naphthalene were able to oxidize 4-chlorophenol to 4-chlorocatechol, or 3-chlorobenzoate to 3-chloro-5,6-dihydroxybenzoate (Figure 4.3) even though they were unable to use them for growth (Spokes and Walker 1974). b. The oxidation of nitrobenzene to 3-nitrocatechol by strains of Pseudomonas sp. grown with toluene or chlorobenzene (Haigler and Spain 1991). c. The oxidation of 2-chloronaphthalene to chloro-2-hydroxy-6-ketohexa-2,4-dienoic acids (Figure 4.4) by cells in which biotransformation of the substrate was induced by growth with naphthalene (Morris and Barnsley 1982).

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CO2H Cl

Cl

Cl

Cl

Cl

Cl

OH

Cl

Cl

Cl

OH

FIGURE 4.2

Cl

OH

OH

OH

Biotransformation of 2,3,6-trichlorobenzoate. OH

CO 2 H

CO 2 H

OH

CO 2H

HO

HO

; Cl Cl

Cl

HO

Cl

Cl

FIGURE 4.3 Biotransformation of 4-chlorophenol and 3-chlorobenzoate. Cl

Cl

CO 2 H

OH CO 2 H CHO

OH Cl

or Cl

OH

CO 2 H

FIGURE 4.4

Cl

CHO CO 2 H OH

Biotransformation of 2-chloronaphthalene.

d. The degradation of polychlorinated biphenyls has been established in a number of organisms enriched with biphenyl as carbon source, for example, Pseudomonas sp. strain LB400 (Bopp 1986), Alcaligenes eutrophus strain H850 (Bedard et al. 1987), and P. pseudoalcaligenes (Furukawa and Arimura 1987). e. The use of brominated biphenyls has been examined to induce anaerobic dechlorination of highly chlorinated biphenyls including the hepta-, hexa-, and pentachloro congeners (Bedard et al. 1998). Di- and tribromo congeners were the most effective and were themselves reduced to biphenyl. f. The oxidation of methylbenzothiophenes by cells grown with 1-methylnaphthalene (Saflic et al. 1992). g. The mineralization of benz[a]anthracene, chrysene, and benzo[a]pyrene by organisms isolated by enrichment with and grown at the expense of phenanthrene (Aitken et al. 1998). Comparable situations could be encountered during experiments on bioremediation in which it may be experimentally expedient to grow cells on a suitable analog (Klecka and Maier 1988), or to introduce the organisms into the contaminated site (Harkness et al. 1993). There are, however, inherent dangers in this procedure. For example, cells able to degrade 5-chlorosalicylate (Crawford et al. 1979), 2,6dichlorotoluene (Vandenbergh et al. 1981) and pentachlorophenol (Stanlake and Finn 1982) were unable to degrade the nonhalogenated analogs. Further, the degradation of 4-nitrobenzoate was inhibited by benzoate even though the strain could use either substrate separately (Haller and Finn 1978).

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ENZYME INDUCTION BY GROWTH ON STRUCTURALLY UNRELATED COMPOUNDS Enzymes necessary for the metabolism of a substrate may be induced by growth on structurally unrelated compounds. In the examples used for illustration, monooxygenases play a cardinal role as a result of the versatility of methane monooxygenase, while monooxygenases that may be involved in toluene degradation are discussed in Chapter 3, Part 1 and Chapter 8, Part 1. 1. Trichloroethene and aromatic compounds. A striking example is the degradation of trichloroethene by different strains of Pseudomonas sp. grown with phenol (Folsom et al. 1990) or with toluene. This capability has attracted attention for the bioremediation of sites contaminated with chloroethenes (Hopkins and McCarty 1995). Conversely, toluene degradation is induced (a) by trichloroethene in a strain of P. putida (Heald and Jenkins 1994) and (b) in P. mendocina (though not in Burkholderia [Pseudomonas] cepacia or P. putida strain F1) by trichloroethene, pentane, and hexane (McClay et al. 1995). This metabolic versatility may be consistent with the different pathways that are followed in toluene degradation (Figure 4.5): a. P. putida F1 by the classical toluene dioxygenase system (Zylstra et al. 1989) b. B. cepacia G4 by monooxygenation to 2-methylphenol (Shields et al. 1989) c. P. mendocina KR by monooxygenation to 4-methylphenol (Whited and Gibson 1991) d. P. pickettii PKO1 by monooxygenation to 4-methylphenol (Olsen et al. 1994) The last example is mediated by a monooxygenase that can be induced by benzene, toluene, and ethylbenzene, and also by xylenes and styrene. A plausibly analogous situation exists for strains of Pseudomonas sp. and Rhodococcus erythropolis that were obtained by enrichment with isopropylbenzene, and that could be shown to oxidize trichloroethene (Dabrock et al. 1992). In addition, one of the pseudomonads could oxidize 1,1-dichloroethene, vinyl chloride, trichloroethane, and 1,2-dichloroethane. 2. Methane, butane, and chloroform. Cells of Methylosinus trichosporium grown with methane and of Pseudomonas butanovora and Mycobacterium vaccae grown with butane were able to partially degrade chloroform (Hamamura et al. 1997). Again this may be the result of the induction of monooxygenase activity.

CH 3

CH 3

CH 3

CH 3

CH 2 OH

H

OH

OH OH

OH

H CO 2 H

FIGURE 4.5 Pathways for the biotransformation of toluene.

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3. Strain G4/PR1 of Burkholderia cepacia in which the synthesis of toluene-2-monooxygenase is constitutive is able to degrade a number of ethers including diethyl ether and n-butyl methyl ether but not t-butyl methyl ether (Hur et al. 1997). 4. A good example is provided by DDE, which is the first metabolite in the conventional degradation pathway of DDT and is apparently persistent in the environment. Pure cultures of aerobic and anaerobic bacteria that were able to degrade 1,1-dichloroethene and 4,4′-dichlorobiphenyl, which were considered to represent important structural features of DDE, were, however, unable to degrade DDE even during incubation with dense cell suspensions (Megharaj et al. 1997). Cell extracts of the aerobic organisms were also ineffective, and it was therefore concluded that recalcitrance lay in the structure of 1,1-diphenyl-2,2-dichloroethene since 1,1-diphenylethene could be used as sole substrate for the growth of styrene-degrading strains of Rhodococcus sp. (Megharaj et al. 1998). In contrast, cells of Pseudomonas acidovorans strain M3GY during growth with biphenyl have been shown to degrade DDE with the fission of one ring and production of 4-chlorobenzoate (Hay and Focht 1998). This situation may be of widespread occurrence and further examples of its existence will be facilitated by insight into the mechanisms of pathways for biodegradation.

ROLE OF READILY DEGRADED SUBSTRATES Results from experiments on biodegradation in which readily degraded substrates such as glucose are added have probably restricted relevance to natural ecosystems in which such substrates exist in negligible concentration. However, readily degraded substrates in addition to those less readily degradable undoubtedly occur in biological-waste-treatment systems. In these circumstances, at least three broadly different metabolic situations may exist: 1. The presence of glucose may suppress degradation of a recalcitrant compound: a. Strains of Pseudomonas pickettii degrade 2,4,6-trichlorophenol, and its degradation is induced by several other chlorophenols, but is repressed in the presence of glucose or succinate (Kiyohara et al. 1992). b. The presence of glucose decreased the rate of degradation of phenol in a natural lake-water community, though the rate was increased by arginine (Rubin and Alexander 1983). 2. The presence of glucose may, however, enhance the degradation of a recalcitrant compound. Several different metabolic situations may be discerned, each probably representing a different mechanism for the stimulation: a. Experiments in which degradation of fluorobenzoates by a mixed bacterial flora was enhanced by the presence of glucose might plausibly be attributed to an increase in the cell density of the appropriate organism(s) (Horvath and Flathman 1976). A comparable conclusion could also be drawn from the data for the degradation of 2,4dichlorophenoxyacetate and O,O-dimethyl-O-[3-methyl-4-nitrophenyl] phosphorothioate in cyclone fermentors (Liu et al. 1981). b. The presence of readily degraded substrates such as glutamate, succinate, or glucose had a stimulatory effect on the degradation of pentachlorophenol by a Flavobacterium sp., and could be attributed to ameliorating the toxic effects of pentachlorophenol (Topp et al. 1988). These additional substrates also enhanced the ability of natural communities to degrade a number of xenobiotics (Shimp and Pfaender 1985a). c. The presence of glucose facilitated the anaerobic dechlorination of pentachlorophenol and may plausibly be attributed to the increased level of reducing equivalents

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(Henriksen et al. 1991). A comparable phenomenon is the enhanced dechlorination of tetrachloroethene in anaerobic microcosms by the addition of carboxylic acids including lactate, propionate, butyrate, and crotonate (Gibson and Sewell 1992). 3. The xenobiotic may be degraded in preference to glucose, which is not a universal growth substrate. This situation is encountered in a phenol-utilizing strain of the yeast Trichosporon cutaneum that possesses a partially constitutive catechol 1,2-dioxygenase (Shoda and Udaka 1980), and illustrates the importance of regulatory mechanisms in determining the degradation of xenobiotics. Constitutive synthesis of the appropriate enzyme systems may indeed be of determinative significance in many natural ecosystems, and merits investigation. In some circumstances, therefore, the presence of readily degraded substrates may clearly facilitate the degradation of recalcitrant xenobiotics, although a generally valid mechanism for these positive effects has not emerged. Whereas the addition of metabolizable analogs may increase the overall rates of degradation (Klecka and Maier 1988), it should be emphasized that the presence of readily degraded substrates in enrichments would generally be expected to be counterselective to the development of organisms that could degrade a specific xenobiotic. In addition, the observed enhancements summarized earlier were generally observed during relatively short time intervals. Simultaneous metabolism of two structurally related substrates in which only one of them serves as growth substrate during biotransformation of the other may exist. A simple example is the O-demethylation of 4,5,6-trichloroguaiacol to 3,4,5-trichlorocatechol followed by successive O-methylation during the growth of a strain of Rhodococcus sp. at the expense of vanillate (Figure 4.6) (Allard et al. 1985). The results of experiments with mixtures of benzoate and 2,5-dichlorobenzoate using variants of a strain of Pseudomonas aeruginosa grown in chemostat cultures have revealed the following important aspects of environmental significance (van der Woude et al. 1995): a. Variants that formed stable cultures could be obtained by growth limitation with both benzoate and 2,5-dichlorobenzoate. b. One of these variants was capable of 2,5-dichlorobenzoate-limited growth at an oxygen concentration of 11 µM, although the presence of benzoate increased the residual concentration of 2,5-dichlorobenzoate from 0.05 to 1.27 µM. A number of factors may be responsible for these observations, including the oxygen gradient within the cell and the oxygen concentration required for synthesis of the degradative oxygenases. It was concluded that under the low oxygen tensions that might exist in natural ecosystems, the presence of the more readily degraded benzoate necessitated high oxygen affinity for organisms to achieve complete degradation of the 2,5-dichlorobenzoate. In this case, it is relevant to note that the organism was originally isolated after enrichment at high oxygen concentrations. The issue of oxygen concentration is discussed in a wider context later in this chapter.

OH Cl

OCH 3

Cl

Cl

OH

Cl

Cl

OCH 3 OH

Cl

Cl Cl

OCH 3

OH

Cl

OCH 3

Cl Cl

Cl

FIGURE 4.6 Biotransformation of 4,5,6-trichloroguaiacol by Rhodococcus sp. during growth with vanillate.

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PHYSICAL PARAMETERS Physical parameters such as temperature, salinity, pH, and oxygen concentration may critically determine the persistence or otherwise of a xenobiotic under natural conditions, and these should therefore be critically evaluated. Laboratory experiments can be carried out under any of the conditions that simulate the natural environment, and these can be imposed during isolation of the organisms by enrichment and incorporated into the design of subsequent experiments on biodegradation and biotransformation. In practice, most experiments are carried out with mesophiles and at pH values in the vicinity of pH 7, presumably motivated by the fact that these are, or are assumed to be, prevalent in natural ecosystems. It should be emphasized that soil and water temperatures during the winter in high latitudes of both the northern and southern hemispheres are low—probably below 10°C. Greater advantage should therefore be taken of psychrophiles, particularly in investigations aimed at providing realistic estimates of rates for microbial transformation under natural conditions. With the notable exception of the anaerobic sulfate-reducing bacteria, surprisingly few investigations have used truly marine bacteria in spite of the fact that substantial quantities of xenobiotics are discharged into the open sea either directly or via input from rivers. These issues are discussed in the next section.

TEMPERATURE The role of temperature may be of particular significance for mixed cultures of organisms in the natural environment, or when they are used in laboratory studies. Temperature may effect important changes both in the composition of the microbial flora and on the rates of reaction. Greatest attention has hitherto been directed to hydrocarbons and PCBs. Some illustrative examples of important issues that are especially relevant to bioremediation include the following: 1. An anaerobic sediment sample was incubated with 2,3,4,6-tetrachlorobiphenyl at various temperatures between 4 and 66°C (Wu et al. 1997a). The main products were 2,4,6- and 2,3,6-trichlorobiphenyl, and 2,6-dichlorobiphenyl. The first was produced maximally and discontinuously at 12 and 34°C, the second maximally at 18°C, and the third was dominant from 25 to 30°C. Dechlorination was not observed above 37°C. In a further comparable study with Arochlor 1260, hexa- to nanochlorinated congeners were dechlorinated with a corresponding increase in tri- and tetrachloro congeners, and four dechlorination patterns associated with different temperature ranges could be distinguished (Wu et al. 1997b). 2. Sediment samples from a contaminated site were spiked with Arochlor 1242 and incubated at 4°C for several months (Williams and May 1997). Degradation by aerobic organisms in the upper layers of the sediment—but not in those at >15 mm from the surface—occurred with the selective production of di- and trichlorobiphenyls. Some congeners, including 2,6- and 4,4-dichlorobiphenyls and a range of trichlorobiphenyls, were not found. 3. The enrichment of arctic soils from Northwest Territories, Canada with biphenyl yielded organisms that were assigned to the genus Pseudomonas (Master and Mohn 1998). Rates of removal of individual congeners of Arochlor 1242 were examined at 7°C, and compared with those for the mesophilic Burkholderia cepacia strain LB400. The spectrum of rates for all congeners was similar for the arctic strains and those for some of the trichlorinated congeners was considerably greater than those obtained for Burkholderia cepacia strain LB400. 4. The mineralization of 14C-ring-labeled toluene was examined in contaminated and uncontaminated samples from an aquifer in Alaska and one in South Carolina (Bradley and Chapelle 1995). A number of important facts emerged: a. Rates for organisms from South Carolina were greater for the contaminated sample than for the uncontaminated sample, and for the uncontaminated sample were zero at 5°C.

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b. Rates for organisms from the contaminated site in Alaska were highly sensitive to temperature, and showed a distinct maximum at 20°C that was not apparent for those from the uncontaminated site. In contrast, rates for both groups of organisms were virtually identical at 5 and 35°C. 5. The aerobic degradation of hydrocarbons has been examined in many cold-temperature regions—Arctic (Whyte et al. 1996; Yu et al. 2000), Alpine (Margesin et al. 2003), and Antarctic (Ruberto et al. 2003; Aislabie et al. 2004)—and in both terrestrial and marine environments. A number of psychotrophic organisms that were isolated from various sites in northern Canada were capable of mineralizing naphthalene, toluene, and linear dodecane and hexadecane at 5°C. Although some of the positive strains possessed genes homologous with those required for metabolism of the substrates by established pathways—nahB for naphthalene and xylE and todC1 for toluene—others showed only low homologies and may have possessed novel catabolic genes (Whyte et al. 1997). Two organisms assigned to the genus Pseudomonas degraded alkanes (C5–C12), and toluene, and naphthalene at both 5°C and 25°C (Whyte et al. 1998). It was shown by PCR and DNA sequence analysis that the plasmid-borne catabolic genes were comparable to those for the alk pathway in P. oleovorans and the nah pathway in P. putida G7. 6. Anaerobic psychrophilic sulfate-reducing bacteria isolated using lactate as substrate were assigned to the genera Desulfofrigus, Desulfofaba, and Desulfotalea, and could grow at temperatures down to –1.8°C. Most strains were able to use Fe(III)citrate as electron acceptors as well as sulfate, and a range of C1–C4 carboxylates as electron donors (Knoblauch et al. 1999). 7. Experiments using a fluidized-bed reactor showed that the removal of chlorophenols could be accomplished by organisms adapted to a temperature of 5–7°C (Järvinen et al. 1994).

OXYGEN CONCENTRATION The question of oxygen concentration is a good deal more complex. The extreme conditions in which oxygen concentration is high, for example in well-mixed surface waters, or essentially absent, for example, in organic-rich deep sediments, are readily visualized. Although these conditions can be realized in laboratory experiments, it is worth drawing attention to a number of less clear-cut situations: 1. Under conditions of oxygen limitation in continuous culture, mixed populations of aerobic bacteria and the strictly anaerobic Methanobacterium formicicum and Methanosarcina barkeri could be maintained (Gerritse and Gottschal 1993a). 2. The existence of microaerophilic organisms such as Wolinella succinogenes that was formerly considered an anaerobe or, conversely, the oxygen tolerance of many clostridia, suggests that such organisms may occupy an ecological niche between the two extremes already noted. 3. Although Pseudomonas aeruginosa is normally considered an aerobic organism, growth can take place by denitrification or by fermentation of arginine. There are also several lines of evidence that support its tolerance of oxygen limitation that may be relevant to its survival in biofilms and it role in pathogenesis: a. Under oxygen stress, Pseudomonas aeruginosa strain PAO1 produces polysaccharide on the cell surface that may limit the diffusion of oxygen into the cell (Sabra et al. 2002). b. Pseudomonas aeruginosa strain PAO1 was able to survive anaerobic conditions by the fermentation of pyruvate. Although pyruvate is incapable of supporting growth, it is transformed under anaerobic conditions to lactate, acetate, and succinate (Eschbach et al. 2004).

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c. During pyruvate fermentation in a complex medium, arginine deiminase (ArcA), catabolic ornithine carbamoyltransferase (ArcB) and carbamate kinase (ArcC) were induced, and survival of the organism was dependent on a protein (PA3309) related to universal stress proteins (Schreiber et al. 2006). Oxygen concentration is important in a range of contexts, for example, (a) regulating the synthesis of peroxidases in white-rot fungi, (b) determining the outcome of the degradation of the trichloromethyl group in a substituted pyridine by Nitrosomonas europaea, and (c) inducing cytochrome P45014DH in Saccharomyces cerevisiae. Anaerobic Bacteria It is well established from laboratory practice that anaerobic microorganisms differ widely in their oxygen tolerance, ranging from the strictly anaerobic methanogens to the more tolerant clostridia. Although strictly anaerobic bacteria do not generally grow in the presence of high-potential electron acceptors such as oxygen or nitrate, there are a number of intriguing exceptions. For example, an obligately anaerobic organism uses nitrate as electron acceptor during the degradation of resorcinol (Gorny et al. 1992). The normally strictly anaerobic sulfate-reducing organisms Desulfobulbus propionicus and Desulfovibrio desulfuricans may grow by reducing nitrate or nitrite to ammonia using hydrogen as the electron donor (Seitz and Cypionka 1986). This draws attention to the unknown extent to which such organisms exist in natural systems, since strictly anaerobic conditions are not always used for the isolation of organisms using nitrate as an electron acceptor. A few examples illustrate the occurrence of oxygen tolerance among bacteria that are normally considered as strictly anaerobic: a. Desulfovibrio vulgaris Hillsborough is capable of growth at oxygen concentrations of 0.24–0.48 µM, and it was suggested that this organism might protect anoxic environments from adverse effects resulting from the intrusion of oxygen (Johnson et al. 1997). b. The anaerobic Geobacter sulfurreducens is able to use oxygen as terminal electron acceptor, and can grow at oxygen concentrations up to 10% (Lin et al. 2004). It was suggested that this would contribute to its survival under oxic conditions. c. Although bacteria belonging to the genus Clostridium are generally considered to be obligate anaerobes, there are important exceptions. Clostridium acetobutylicum and Cl. aminovalericum can grow under microaerophilic conditions and consume oxygen. The gene that encoded NADH oxidase was unregulated in Cl. aminovalericum when microaerophilic conditions were attained, and a number of NAD(P)H-dependent enzyme activities were induced in Cl. acetobutylicum (Kawasaki et al. 2005). d. Among bacteria considered strict anaerobes, cytochrome bd-oxidases have been found, and these enable the organisms to tolerate and reduce low levels of oxygen. Such cytochromes have been found in Bacteroides fragilis (Baughn and Malamy 2004), Desulfovibrio gigas (Lemos et al. 2001), and Moorella thermoacetica (Das et al. 2005). Aerobic and Facultatively Anaerobic Organisms 1. Facultatively anaerobic organisms can employ either fermentative or oxidative modes for the metabolism of appropriate substrates such as carbohydrates, while others can use nitrate as an alternative electron acceptor in the absence of oxygen. Such organisms have, therefore, two metabolic options, even though they may often be mutually exclusive. An illustrative example of the metabolic flexibility of facultatively anaerobic organisms is provided by the type strains of all the species of the enteric genus Citrobacter. These organisms were able to degrade a number of amino acids including glutamate using either respiratory or fermentative metabolism, and they could rapidly switch between these alternatives (Gerritse and Gottschal 1993b). Under anaerobic conditions, the initial steps of

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glutamate degradation involved the formation of 3-methylaspartate, mesaconate, and citramalate, which are typical of clostridial fermentations. Although the oxygen concentration in natural environments may be highly variable, organisms with such a high degree of metabolic flexibility may reasonably be presumed to be at an advantage. 2. Oxygen toxicity under aerobic conditions may result from the synthesis of toxic compounds including superoxide, hydrogen peroxide, or hydroxyl radicals. Further details have emerged from observations that the Fe isoenzyme of superoxidase dismutase is synthesized by Escherichia coli under anaerobic conditions, in contrast to the Mn isoenzyme that is produced only under aerobic conditions. The synthesis of the former facilitates the transition from anaerobic to aerobic conditions by destroying the superoxide radical generated after exposure to oxygen (Kargalioglu and Imlay 1994). This flexibility is important in environments where fluctuating oxygen concentrations prevail. Studies with mutants of Pseudomonas aeruginosa in which the synthesis of Mn superoxide dismutase and Fe superoxide dismutase were impaired showed that growth in a complex medium or in a defined glucose medium was adversely affected by the latter but only insignificantly by the former (Hassett et al. 1995). Concentrations of Fe must be regulated, since excess of Fe within the cell would result in the production of deleterious hydroxyl radicals from molecular oxygen (Touati et al. 1995). The transport of Fe into the cell is regulated by Fe(III) high-affinity siderophores which are synthesized within the cell and excreted into the medium before transport of the complexes into the cell by specific membrane transporters (Wandersman and Delepelaire 2004). In this way, homeostasis is achieved. 3. There is substantial evidence that organisms that are strictly dependent on the aerobic metabolism of substrates for growth and replication may nonetheless accomplish biodegradations and biotransformations under conditions of low oxygen concentration. Indeed, such conditions may inadvertently prevail in laboratory experiments using dense cell suspensions; it is important to appreciate that the growth of these aerobic organisms may be limited by the availability of oxygen. A number of examples are given below to illustrate the environmental role of putatively aerobic organisms under conditions of low oxygen concentration: a. The rate of biodegradation of hexadecane in a marine enrichment culture was unaffected until oxygen concentrations were lower than 1% saturation (Michaelsen et al. 1992). b. It has been shown that the degradation of pyrene by a strain of Mycobacterium sp. can take place at low oxygen concentrations (Fritzsche 1994). c. Alcaligenes sp. strain L6 that was obtained after enrichment with 3-chlorobenzoate in an atmosphere of only 2% oxygen in the gas phase possessed both 2,5- and 3,4dihydroxybenzote dioxygenase activities although it lacked catechol dioxygenase activity that would be required for fission of the aromatic ring (Krooneman et al. 1996). The affinity for the substrate of cells grown under oxygen limitation was three times greater than was observed under excess oxygen, and exceeded values for other bacteria growing with benzoate or 2,5-dichlorobenzoate. This strain that metabolized 3-chlorobenzoate by the gentisate pathway was, under low oxygen concentrations, able to compete successfully with a Pseudomonas sp. strain A that used the catechol pathway (Krooneman et al. 1998). d. Further details emerged from the results of experiments that compared the kinetic properties of catechol-2,3-dioxygenases from toluene-degrading pseudomonads isolated from aquifer sands or groundwater (strains W 31 and CFS 215) with other strains (P. putida F1). This revealed a number of significant differences that may plausibly be associated with the low oxygen concentrations pertaining in the environment from which the strains were isolated. For strains that degraded toluene via cis-toluene2,3-dihydrodiol and 3-methylcatechol, lower values of Km were observed for O2 and 3-methylcatechol, whereas higher values of Vmax for O2 and 3-methylcatechol were found (Kukor and Olsen 1996) (Table 4.1).

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TABLE 4.1 Values of Km and Vmax for O2 and 3-Methylcatechol in Strains F1, W31, and CFS 215 Vmax Substrate

Km Substrate Strain

O2

3-Methylcatechol

O2

3-Methylcatechol

F1 W31 CFS 215

9.7 2.0 0.9

16.9 0.3 0.5

4.3 293 400

17 125 180

e. The growth of Pseudomonas putida strains KT2442 and mt-2 on aromatic carboxylates decreased with O2 concentrations <10 µM and under these conditions strain KT2442 excreted catechol during growth with benzoate, or 3,4-dihydroxybenzoate from 4hydroxybenzoate (Arras et al. 1998). In these experiments, the rate of oxygenation rather than the pO2 was the limiting factor, since the Km for catechol 1,2-dioxygenase was 20 µM O2. f. The synthesis of cytochrome P450 in yeasts that assimilate long-chain alkanes has been examined in facultatively anaerobic species of Candida. In Candida guilliermondii, levels of cytochrome P450 that were induced during growth with hexadecane—though not with glucose—increased at dissolved O2 concentrations <20% (Mauresberger et al. 1980). For cells of Candida tropicalis grown with hexadecane, levels of cytochrome P450 increased when the oxygen concentration was <2kPa (14% dissolved O2). However, the levels of alcohol dehydrogenase, aldehyde dehydrogenase, catalase, and cytochrome c were not influenced (Gmünder et al. 1981). 4. An issue of particular relevance in the context of bioremediation is illustrated by the increased rate of cell death in an established naphthalene-degrading Pseudomonas putida G7 brought about by the substrate (naphthalene) under conditions of oxygen (or combined oxygen and nitrogen) limitation (Ahn et al. 1998). 5. The growth of Bacillus subtilis may take place under a variety of conditions: (a) aerobic conditions, (b) using nitrate as electron acceptor, and (c) fermentative conditions with glucose provided pyruvate is available as an electron acceptor since the organism lacks pyruvate formate hydrogen lyase (Nakano and Zuber 1998). 6. Thauera sp. strain DNT-1 is able to degrade toluene under aerobic conditions mediated by a dioxygenase, and under denitrifying conditions in the absence of oxygen by the anaerobic benzylsuccinate pathway (Shinoda et al. 2004). Whereas the tod genes were induced under aerobic conditions, the bss genes were induced under both aerobic and anaerobic conditions. Strains may be able to regulate the pathway of degradation of toluene to the availability of oxygen: aerobic degradation by dioxygenation and under denitrifying conditions in the absence of oxygen by the benzylsuccinate pathway. It should also be appreciated that the oxygen concentration may determine the outcome of a reaction, or even the possibility of either oxidative or reductive reactions: a. The degradation of pentachlorophenol by Mycobacterium chlorophenolicum (Rhodococcus chlorophenolicus) proceeds by initial formation of tetrachlorohydroquinone. Whereas its formation is oxygen-dependent, its subsequent degradation can be accomplished in the absence of oxygen (Apajalahti and Salkinoja-Salonen 1987). b. Pseudomonas putida G786 containing the CAM plasmid carried our reductive dehalogenation of a number of haloalkanes (Hur et al. 1994) and when a plasmid containing

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HO 2C

N

Cl

Cl2 CH

N

Cl

(a)

Cl3 C

FIGURE 4.7

N

Cl

(b)

Biotransformations of 2-chloro-6-trichloromethylpyridine.

genes encoding toluene dioxygenase was incorporated, both reduction and oxidation were observed (Wackett 1994). c. Biotransformation as opposed to biodegradation may, in fact, be favored by limited oxygen concentration. A good example is provided by the synthesis of 7α,12β-dihydroxy-1,4androstadien-3,17-dione from cholic acid by a strain of Pseudomonas sp. (Smith and Park 1984), in which oxygen limitation restricts the rate of C-9 α-hydroxylation that initiates degradation. d. Oxidation of the trichloromethyl group in 2-chloro-6-trichloromethylpyridine to the corresponding carboxylic acid by Nitrosomonas europaea occurs at high oxygen concentrations during cooxidation of ammonia or hydrazine (Figure 4.7a). In contrast, at low oxygen concentrations in the presence of hydrazine, reductive dechlorination to 2-chloro6-dichloromethylpyridine occurs (Vannelli and Hooper 1992) (Figure 4.7b).

REDOX POTENTIAL For strictly anaerobic bacteria, it is usually not sufficient that the medium is anoxic. It must also be poised at the correct redox potential. It is for this reason that prereduced media are used. This is accomplished by the addition of, for example, sulfide, dithionite, or titanium(III) citrate, and the media generally contain a redox indicator such as resazurin. Examples have been given earlier of reactions carried out by aerobic organisms at low oxygen concentrations, and the outcome of reactions carried out by such organisms may be influenced by the redox potential. For example, cells of the aerobic Pseudomonas cepacia carried out the degradation of tetrachloromethane only when a negative potential was maintained, and the maximum rate occurred at a potential of ca. –150 mV (Jin and Englande 1997).

ASSOCIATION OF BACTERIA WITH PARTICULATE MATERIAL: “FREE” AND “BOUND” SUBSTRATES Xenobiotics exist not only in the “free” state but also in association with organic and mineral components of particles in the water mass, and the soil and sediment phases. This association is a central determinant of the persistence of xenobiotics in the environment, since the extent to which the reactions are reversible is generally unknown. Such residues may therefore be inaccessible to microbial attack and apparently persistent. This is a critical factor in determining the effectiveness of bioremediation (Harkness et al. 1993). Although the most persuasive evidence for the significance of reduced bioavailability comes from data on the persistence of agrochemicals in terrestrial systems (Calderbank 1989), the principles can be translated with modification to aquatic and sediment phases that contain organic matter that resembles structurally that of soils.

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BIOLOGICAL MECHANISMS FOR ASSOCIATION WITH ORGANIC COMPONENTS OF SOIL AND SEDIMENT These may involve both the original compound and its metabolites produced by biological reactions. This mechanism has wide implications, and has been most extensively documented in the terrestrial environment. 1. Naphth-1-ol is an established fungal metabolite of naphthalene and may play a role in the association of naphthalene with humic material (Burgos et al. 1996). 2. 13C-labeled metabolites of 9-[13C])-anthracene including 2-hydroxyanthracene-3carboxylate and phthalate were not extractable with acetone or dichloromethane, but could be recovered after alkaline hydrolysis (Richnow et al. 1998). 3. The results of experiments with 14C-labeled pyrene added to a pristine forest soil illustrated a number of important issues (Guthrie and Pfaender 1998): a. Extensive mineralization took place only in samples amended with a pyrene-degrading microbial community. b. Compared with an azide-treated control, there was a substantially greater nonextractable fraction of the label in soils containing either the natural or the introduced microflora. c. Metabolites that could be released by acid and base extraction remained in the soil, even after a 270-d incubation. 4. The metabolism of 14C-labeled BTX has been examined in soil cultures, and a mass balance constructed after 4 weeks of aerobic incubation (Tsao et al. 1998). Mineralization of all substrates was ca. 70% but ca. 20% of the label in toluene and ca. 30% in o-xylene were found in humus. It was suggested that alkylated catechol metabolites were responsible for this association. 5. The mechanism for the interaction of cyprodinil (4-cyclopropyl-6-methyl-2-phenylaminopyrimidine) with soil organic matter has been examined. The association with soil organic carbon was biologically mediated, and it was shown that this increased during incubation for up to 180 d (Dec et al. 1997a). After 169 d of incubation, the fractions obtained by methanol extraction, and the humic acid and fulvic acid fractions after alkali extraction were examined by 13C NMR (Dec et al. 1997b). Both the phenyl and the pyrimidine rings were associated with humic material, though only partly in the form of intact cyprodinil. 6. Considerable attention has been directed to enzymatic reactions mediated by fungal oxidoreductase enzymes such as phenol oxidase, peroxidase, and laccase. These systems have been used to copolymerize structurally diverse xenobiotics to lignin-like structures, and include substituted anilines (Bollag et al. 1983) and benzo[a]pyrene quinone (Trenck and Sandermann 1981). One substantial advantage of using these model systems is that it is possible to isolate the products of the reactions and determine their chemical structures. Some examples are given below to illustrate the different substrates involved and the types of products that may be produced: a. Incubation of pentachlorophenol with a crude supernatant from Phanerochaete chrysosporium in the presence of a lignin precursor (ferulic acid) and H 2O2 produced a high molecular mass polymer (Rüttimann-Johnson and Lamar 1996). It was suggested that this could mimic the association of pentachlorophenol with humic material and the formation of heteropolymers between pentachlorophenol and lignin monomers. b. The reaction between halogenated phenols and syringic acid in the presence of laccase from the fungus Rhizoctonia praticola resulted in the formation of a series of diphenyl ethers. One ring originated from the chlorophenol, together with 1,2-quinonoid products resulting from partial O-demethylation and oxidation (Bollag and Liu 1985) (Figure 4.8). Comparable reactions have also been postulated to occur between 2,4dichlorophenol and fulvic acid (Sarkar et al. 1988).

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207

CO 2H

OCH 3

Cl

Cl

O

Cl Cl

+ Cl Cl

OH + Cl

O

O

O

OCH 3

CH 3O

Cl

OH

OCH 3

Cl

OCH 3

FIGURE 4.8 Reaction between 2,4,5-trichlorophenol and syringic acid catalyzed by laccase. Cl OH

O CH3 O

CH3 O

NH

OH

N Cl

OH

Cl

N

Cl

CH3O

CH3 O O

OH

OCH3 O

FIGURE 4.9 Products from the enzymatic copolymerization of guaiacol and 4-chloroaniline.

c. It has been shown that oligomerization of 4-chloroaniline mediated by oxidoreductases may produce 4,4′-dichloroazobenzene and 4-chloro-4′-aminodiphenyl as well as trimers and tetramers (Simmons et al. 1987). A study with guaiacol and 4-chloroaniline using a number of oxidoreductases demonstrated the synthesis of oligomeric quinone imines together with compounds resulting from the reaction of the aniline with diphenoquinones produced from guaiacol (Simmons et al. 1989) (Figure 4.9). d. Direct evidence for the existence of covalent bonding between 2,4-dichlorophenol and peat humic acid in the presence of horseradish peroxidase has been provided from the results of an NMR study using 2,6-[13C]-2,4-dichlorophenol (Hatcher et al. 1993). In the absence of suitable model compounds, interpretation of the results was based on estimated chemical shifts for a range of plausible structures. The most important contributions came from those with an ester linkage with the phenol group, and covalent bonds between carbon atoms of the humic acid and C-4 (with loss of chlorine) and C-6 of the chlorophenol. e. Laccase-catalyzed reactions between bentazon (3-isopropyl-H-2,1,3-benzothiadiazine4(3H)-one 2,2-dioxide) and various humic acid monomeric components have been studied, and the products from reactions with catechol examined in detail using both 1H and 13C NMR (Kim et al. 1997). Products with masses of 348 and 586 were isolated, and these were assigned to products formed by reactions between the nitrogen atom of bendazon and the 1,2-quinone produced by the laccase. f. Coniferyl alcohol—the monomeric precursor of lignin—was copolymerized by peroxidase and H2O2 in the presence of 15N aniline and 3,4-dichloroaniline in various ratios, and the products were examined by 1H, 13C, and 15N NMR (Lange et al. 1998). The conjugates were formed by reaction at the benzylic carbon atom of the coniferyl alcohol polymer. Although the anilines could be recovered by acid hydrolysis, it was pointed out that this could result from the high molar ratio of anilines used for copolymerization.

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There is therefore extensive evidence that may be used to rationalize the occurrence of “bound” residues in soils, and this phenomenon is of particular significance for agrochemicals. Such processes influence not only their recovery by chemical procedures, but also their biological effect and their biodegradability (Calderbank 1989). The extent to which these principles are applicable to aquatic systems appears to have been established less frequently though it is plausible that comparable mechanisms exist in the environment.

AGING This process should be considered in the light of the preceding comments on association. Many experiments on the recoverability, persistence, and toxicity of xenobiotics have used spiked samples that do not take into account the cardinal issue of alterations in the contaminant that have taken place after deposition. This is termed aging, and should be evaluated critically in determining persistence. Some examples are given below as illustration for both terrestrial and aquatic systems: 1. Experiments have examined the effect of aging on biodegradability using sterilized samples of soils that were spiked and aged under laboratory conditions. These have shown that the rates and extent of degradation of phenanthrene and 4-nitrophenol by an added strain of Pseudomonas sp. decreased markedly with prolonged aging (Hatzinger and Alexander 1995). 2. The sorption–desorption of PAHs has been extensively investigated, and the role of desorption in determining their biodegradability in aged sediments has been widely accepted (references in Carmichael et al. 1997). A definitive study using 14C-phenanthrene and 14 C-chrysene showed that in contaminated soils, their rates of mineralization were much lower than the rates of desorption from spiked sediments. In contrast, for aged substrates desorption rates were essentially comparable to rates of mineralization. This suggested that the indigenous microflora may have become adapted to the low substrate concentrations available by desorption (Carmichael et al. 1997). 3. Suspensions of 2,4-dichlorophenoxyacetate sorbed onto sterile soils were completely protected from degradation by either free or sorbed bacteria, and degradation of the substrate required access of the bacteria to the free compound in solution (Ogram et al. 1985). Rates of degradation in soil with a high organic content were lower than for one with a low organic content (Greer and Shelton 1992), and this adds additional support for the significance of desorption of the xenobiotic in determining its biodegradability. 4. 14C-labeled 2,4-dichlorophenol bound to synthetic or natural humic acids or polymerized by H2O2 and peroxidase was mineralized to CO2 only to a limited extent (<10%), and the greater part remained bound to the polymers (Dec et al. 1990). 5. 14C-labeled 3,4-dichloroaniline-lignin conjugates were degraded to 14CO2 by Phanerochaete chrysosporium as effectively as the free compound (Arjmand and Sandermann 1985), and it was therefore concluded that these “bound” residues were not persistent in the environment. This may, however, represent a special case for the following reasons: (a) although this organism is able to degrade lignin, the relevance of such organisms in most aquatic environments is possibly marginal, (b) the lignin peroxidases implicated in lignin degradation are generally extracellular so that soluble substrates are probably not necessary. 6. The presence of humic acids had a detrimental effect on the degradation of substituted phenols by a microbial community after lengthy adaptation to the humic acids, and was not alleviated by the addition of inorganic nutrients (Shimp and Pfaender 1985b). The diminished number of organisms with degradative capability was responsible for the reduced degree of degradation, so that the predominant effect was probably the toxicity of the humic acids even toward adapted microorganisms.

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7. In short-term experiments with carbofuran (2,3-dihydro-2,2-dimethyl-7-benzofuranyl-Nmethylcarbamate), degradation was accomplished by organisms in an enrichment culture obtained from soils with a low carbon content where sorption of the substrate is low, though it was essentially absent in cultures obtained from soils with a high organic matter (Singh and Sethunathan 1992). 8. Experiments examined a chlorocatechol-contaminated sediment, and interstitial water prepared from it. These showed that the concentrations of total 3,4,5-tri- and tetrachlorocatechols (i.e., including the fraction that is released only after alkaline extraction) were apparently unaltered during prolonged incubation even after addition of cultures with established dechlorinating capability for the soluble chlorocatechols (Allard et al. 1994). Fewer controlled experiments have been carried out for purely aquatic systems. Montmorillonite complexes with benzylamine at concentrations below 200 µg/L decreased the extent of mineralization in lake-water samples, although a similar effect was not noted with benzoate (Subba-Rao and Alexander 1982). Even in apparently simple systems, general conclusions cannot therefore be drawn even for two structurally similar aromatic compounds, both of which are readily degradable under normal circumstances in the dissolved state. All the preceding investigations have been concerned with polar compounds for which plausible mechanisms for their association with organic components of water, soil, and sediment may be more readily conceptualized. To provide a wider perspective, examples are given below for neutral compounds: 1. The aerobic mineralization of α-hexachloro[aaaaee]cyclohexane by endemic bacteria in the soil is limited by the rate of its desorption and by intraparticle mass transfer (Rijnaarts et al. 1990). 2. Whereas degradation of the readily extractable toluene in spiked soil by Pseudomonas putida was rapidly accomplished, there was a reside that was degraded much more slowly at a rate that was apparently dependent on its desorption (Robinson et al. 1990). 3. The extent of bioremediation of sediments contaminated with PCBs appears to be limited by the association of a significant fraction with organic components of the sediment phase (Harkness et al. 1993). 4. Immobilization of neutral xenobiotics in soils by quaternary ammonium cations has been established, and its significance on the bioavailability of naphthalene to bacteria has been examined. Bioavailability was determined by the rates of desorption, and these differed between a strain of Pseudomonas putida and one of Alcaligenes sp. (Crocker et al. 1995). An important development has been the isolation of bacteria that were able to degrade phenanthrene that was sorbed to humic acid material (Vacca et al 2005). Enrichment was carried out with PAH-contaminated soils using phenanthrene sorbed to commercial humic acid. Only the strains isolated from this enrichment were able to carry out degradation of 14C-labeled phenanthrene, and this exceeded by factors of 4–9 the amount estimated to be available from the aqueous phase alone. It was suggested that specially adapted bacteria might interact specifically with naturally occurring colloidal material. The results of these experiments in both aquatic systems and terrestrial systems may profitably be viewed against the extensive evidence for the persistence of agrochemicals in the terrestrial environment. Considerable effort has been directed to the issue of bound residues of agrochemicals (Calderbank 1989) and to its significance in determining both their biological effects and their persistence. This is now fully accepted in contemporary thinking. At the same time, it should be appreciated that from an economic point of view, enhanced rates of degradation of agrochemicals in the terrestrial environment may be highly undesirable (Racke and Coats 1990). Nonetheless, care should be exercised in making generalizations. For example, whereas it has been established that

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soil microorganisms may significantly increase the evolution of 14CO2 from 14C parathion (Racke and Lichtenstein 1985), this was not observed with chlorpyrifos (Racke et al. 1990). The analysis of these diverse observations clearly demonstrates that the persistence of a xenobiotic in the aquatic or in the terrestrial environment may be significantly greater when it is bound either to inorganic minerals or to any of a range of complex natural polymers such as humic and fulvic acids. One of the key issues is the rate of desorption of a xenobiotic from the matrices. This may depend critically on the mechanism of the association, and the extent of bioavailability may also depend on the organisms. For example, (a) soil-sorbed naphthalene was degraded at markedly different rates by two naphthalene-degrading organisms (Guerin and Boyd 1992), and (b) at low substrate concentrations, 2,4-dichlorophenoxyacetate was degraded at different rates by the two strains that were examined (Greer and Shelton 1992). Although a number of important unresolved issues remain, it is clear that the degree of bioavailability of xenobiotics in natural systems introduces an important additional uncertainty in extrapolating to processes and rates in natural ecosystems the results of studies on biodegradation and biotransformation of “free” xenobiotics in laboratory experiments. This is important not only in field investigations. Even in laboratory experiments on the metabolism of xenobiotics, problems of association between the substrate and the microbial cells may occur. If this were not quantitatively evaluated or eliminated, the results and interpretation of such experiments would be seriously compromised. Increasing attention has been directed to the degradation of xenobiotics in aquifers, and it has been shown that most of the relevant bacteria are associated with fine particles rather than existing as free entities. It is therefore important to include such material in laboratory experiments using unenriched communities that attempt to simulate natural conditions (Holm et al. 1992). The interdependence of surfactant sorption and biodegradability supported by the results of laboratory experiments. For example, Pseudomonas sp. strain DES1 that was able to degrade sodium dodecyl sulfate attached to sediment particles did so more effectively than organisms that were unable to degrade analogous (nondegradable) substrates (Marchesi et al. 1994). Although this was attributed to the effect of the metabolite dodecan-1-ol, further study has revealed a more complex situation (Owen et al. 1998) as a result of differences in the responses of different strains and toward different surfactants. Bacterial associations with particulate material should also be evaluated in the context of the bioavailability of the substrate and in bioremediation strategies.

SUBSTRATE CONCENTRATION, TRANSPORT INTO CELLS, AND TOXICITY Procedures used for growth of organisms to be used in studies of biodegradation may be pragmatically divided into major groups: 1. Complex media for organisms that have undetermined and complex nutritional requirements, including, some fungi and anaerobic bacteria 2. Defined mineral media containing only the substance whose biodegradation is being examined 3. Use of analogs of the substrate, for example, biphenyl for PCB-degrading organisms 4. Substrates that are related to those naturally available to the degrading organism, but unrelated to that of the contaminant has been exploited only occasionally: a. The use of a range of coumarin (esculetin) and flavanone (quercitin, naringenin) glycones and aglycones, and a flavane (catechin) (Figure 4.10) as substrates for the anaerobic dechlorination of chlorocatechols (Allard et al. 1991, 1992) b. The growth of established PCB-degrading bacteria on a range of flavones, flavanols, and flavanones (catechin, naringenin, and myricitin) (Figure 4.10) (Donnelly et al. 1994) c. The induction by terpenoids including carvone and limonene (Figure 4.10) of the degradation of PCB congeners (Gilbert and Crowley 1997)

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Determinants and Interactions

211 OH

O

HO

OH

HO

O

O

HO

O

Esculetin

Quercitin

OH OH

OH

OH

O

O

OH

HO

OH

O

HO

Myricitin

O Naringenin

OH

OH

OH OH OH

HO

O

O Catechin

OH

Carvone

Limonene

OH

FIGURE 4.10

Structures of natural products used as substrates or inducers.

UTILIZATION OF LOW SUBSTRATE CONCENTRATIONS Most laboratory experiments on biodegradation and biotransformation have been carried out using relatively high concentrations of the appropriate substrates, even though these may be far in excess of those that are likely to be encountered in natural ecosystems (Subba-Rao et al. 1982; Alexander 1985). This limitation is particularly severe in conventional tests for biodegradability, and seriously restricts the degree to which the results of these experiments are environmentally relevant. The investigations by Nyholm et al. (1992) have addressed this important issue in the marine environment by using 14C-labeled substrates. Except in the immediate neighborhood of industrial discharge, low concentrations of xenobiotics are almost certainly the rule rather than the exception in natural ecosystems, and a number of significant experimental observations should be taken into consideration. Investigations of the flora of natural waters have revealed the presence of bacteria able to grow with extremely low substrate concentrations, which have been termed oligotrophs (Poindexter 1981). It is, of course, well established that organisms such as Aeromonas hydrophila (Van der Kooij et al. 1980), Pseudomonas aeruginosa (Van der Kooij et al. 1982), and a species of Spirillum (Van der Kooij and Hijnen 1984) may proliferate in natural waters supplemented with low concentrations of additional organic carbon. It has been suggested that there exist bacteria specially adapted to such conditions, although doubt has been expressed on the absolute distinction between eutrophic and oligotrophic bacteria (Martin and MacLeod 1984). This distinction is blurred by the fact that after prolonged nutrient starvation under laboratory conditions, initially oligotrophic marine bacteria can be isolated on media containing high substrate concentrations (Schut et al. 1993); these organisms are therefore facultatively oligotrophic. These authors point out that obligate oligotrophy may be determined by life history rather than by invariant physiological characteristics. One critical

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issue seems to be the effectiveness and regulation of substrate transport into the cells (Schut et al. 1995). The metabolism of series of 1,2,4,5-tetra-, 1,2,4-tri-, and three isomeric dichlorobenzenes was examined in Burkholderia sp. strain PS14. At nanomolar concentrations in liquid culture, all of them were rapidly metabolized from initial concentrations of 500 nM to below their detection limits, whereas in soil microcosms tetrachlorobenzene at 65 ppb or trichlorobenzene at 54 ppb was metabolized only after 72 h (Rapp and Timmis 1999). An important development that improved the range of bacteria that could be isolated from soil samples in pure culture has been described by Janssen et al. (2002). They used a medium containing 80 mg/L nutrient broth for isolation and succeeded in obtaining representatives of the divisions Actinobacteria, Acidobacteria, and Verrucomicrobia, which were identified on the basis of matching 16S rRNA gene sequences. An impressively high rate of isolation was achieved, although the reasons for success have not been finally resolved; these could include the lengthy incubation, the use of gellan gum in place of agar for solid media, or the addition of CaCl2. This procedure made it possible to isolate the first pure culture of Chthoniobacter flavus (Sangwan et al. 2004). Additional experiments using the V55 medium with a range of carbon sources at 0.5 mM concentration and incubation times up to 2 months resulted in extension of the range of bacteria belonging to the phyla Acidobacteria, Actinobacteria, and Proteobacteria (Davis et al. 2005).

EXISTENCE OF THRESHOLD CONCENTRATIONS This has received only fragmentary attention, but is relevant when bacteria are exposed to extremely low concentrations of a xenobiotic—of the order of nanograms per liter or less. For example, although the rates of biodegradation of phenol, benzoate, benzylamine, 4-nitrophenol, and di (2-ethylhexyl)phthalate in natural lake water were linear over a wide range of substrate concentrations between nanograms per liter and micrograms per liter (Rubin et al. 1982), it has been shown that the rates of degradation of 2,4-dichlorophenoxyacetate at concentrations of the order of micrograms per liter were extremely low (Boethling and Alexander 1979). This observation has subsequently been extended to a greater range of compounds (Hoover et al. 1986). These and other data could be interpreted as supporting the concept of a threshold concentration below which growth and degradation does not take place—or occurs at insignificant rates. Although the reasons for the existence of such threshold concentrations have not been entirely resolved, a number of plausible hypotheses may be put forward: 1. The substrate concentrations may be too low for effective transport into the cells. 2. There may be a limiting substrate concentration required for induction of the appropriate catabolic enzyme. At low substrate concentrations the necessary enzymes would simply not be synthesized, and this could be the determining factor in some circumstances (Janke 1987). Two contrasting results are instructive. Experiments with chlorinated benzenes in which the effect of substrate concentration was examined in batch cultures and in recirculating fermentors showed that although substrates could be degraded completely in the former, a residual concentration of the substrate persisted in the latter (Van der Meer et al. 1992). However, experiments using Burkholderia sp. strain PS14 failed to detect residual concentrations after mineralization of 1,2,4,5tetra- and 1,2,4-trichlorobenzene at concentrations > 0.5 nM (Rapp and Timmis 1999). These results are highly relevant to the issue of bioremediation, which is discussed in Chapters 12 through 14. All these observations emphasize that tests for biodegradability carried out at high substrate concentrations may not adequately predict the rates of degradation occurring in natural ecosystems where only low concentrations of xenobiotics are encountered (Alexander 1985). This phenomenon is therefore of enormous environmental importance, since it would imply the possible extreme persistence of low concentrations in natural ecosystems. The further exploration of this phenomenon is probably only limited in practice by the access to analytical methods for measuring sufficiently

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accurately substrate concentrations at the level of nanograms per liter or lower. Most studies have been carried out using 14C-labeled substrates, which has restricted the range of accessible compounds, although the use of 13C-labeled substrates and application of NMR could extend the range of possibilities.

STRATEGIES USED BY CELLS FOR SUBSTRATES WITH LOW OR NEGLIGIBLE WATER SOLUBILITY This issue is critical to the design of laboratory experiments on biodegradation and biotransformation. Since many contaminants such as alkanes, PAHs, and polyhalogenated compounds including PCBs belong to this class of compounds, water solubility may present a serious obstacle in attempts at the bioremediation of contaminated terrestrial systems. Indigenous organisms may have been able to develop their own strategies to circumvent low water solubility of their substrates, and a few examples of mechanisms are given below as illustration: 1. They may produce extracellular enzymes, which attack the substrate without the need for transport into the cell, for example, cellulase, DNAse, or gelatinase. 2. They may synthesize surface-active emulsifying compounds during growth. This problem has been extensively investigated on account of its commercial application (Gerson and Zajic 1979), and details of the structural aspects of biosurfactants elaborated for the degradation of water-insoluble substrates have been given (Hommel 1994). Some of the key conclusions from a wide range of studies are briefly summarized: a. Glycolipids consisting of long-chain carboxylic acids and rhamnose (Itoh and Suzuki 1972; Rendell et al. 1990) or trehalose (Suzuki et al. 1969; Singer et al. 1990) have been isolated during the growth of a number of different bacteria on n-alkanes. The rhamnolipid surfactant produced by a strain of Pseudomonas sp. was effective in enhancing degradation of octadecane (Zhang and Miller 1992), though the concentration used in these experiments was rather high (300 mg/L) to encourage its practical application. A model that included the effect of rhamnolipids on solubilization, biodegradation, and bioavailability within surfactant micelles has been presented to rationalize the data from batch studies on the dissolution, bioavailability, and biodegradation of phenanthrene (Zhang et al. 1997). b. A polyanionic heteropolysaccharide (emulsan) is produced during the growth of a strain of Acinetobacter calcoaceticus with hydrocarbon mixtures, and the high-molecularweight polymer is necessary for emulsifying activity (Shoham and Rosenberg 1983). The value of emulsan for treating oil spills seems, however, equivocal in the light of results that demonstrate reduced biodegradation in its presence (Foght et al. 1989). c. A Rhodococcus sp. synthesized a glycolipid during growth with n-alkanes and n-alkanols though not with carboxylic acids, triglycerides, or carbohydrates, and its formation was favored by nitrogen limitation (Singer and Finnerty 1990). d. Pseudomonas maltophilia produced an extracellular surfactant during growth with naphth-1-oate, and displayed much greater emulsifying activity toward aromatic hydrocarbons with one- or two-rings than toward aliphatic hydrocarbons (Phale et al. 1995). e. A strain of Pseudomonas sp. produced a surfactant in the presence of high concentrations of glucose or mannitol, and naphthalene or phenanthrene (Déziel et al. 1996). f. The synthesis of an emulsifying agent produced by Candida lipolytica is inducible during growth with a number of n-alkanes, but it is not synthesized during growth with glucose (Cirigliano and Carman 1984). Different mechanisms have therefore clearly emerged and it seems premature to draw general conclusions especially in the application of synthetic and natural surfactants to bioremediation, which is discussed in greater detail in Chapter 14. It is important to note, however, that the production of biosurfactants may not be the only mechanism for facilitating the uptake of substrates with

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low water solubility. For a strain of Rhodococcus sp. that did not produce surfactants, the rates of degradation of pyrene dissolved in water or in the nondegradable 2,2,4,4,6,8,8-heptamethylnonane exceeded those predicted for physicochemical transfer from the solvent to the aqueous phase. They could, however, be accounted for on the basis of uptake both from the phase interface, and from the aqueous solution (Bouchez et al. 1997).

TRANSPORT MECHANISMS The mechanisms whereby carbohydrates, carboxylic acids, and glycerol are transported across the bacterial cell membrane prior to metabolism have been elucidated in great detail. Attention should, however, be drawn to an investigation of the regulation of the transport of amino acids and glucose into a facultatively oligotrophic marine bacterium that revealed the lack of specificity of the constitutive system for amino acids, and the interaction of the regulatory systems for amino acids and glucose (Schut et al. 1995). It should also be noted that different transport systems for C4 dicarboxylic acids into Escherichia coli operate under aerobic and anaerobic conditions (Six et al. 1994), so that uptake is dependent on the physiological state of the cells. In contrast, less effort has apparently been directed to the transport of xenobiotics, and there is an intrinsic difficulty that in contrast to organisms that utilize carbohydrates or amino acids, suitable mutants defective in the metabolism of the substrate may not be available. This limitation makes it impossible to determine directly whether active transport is involved. Although the genes encoding permeases have been described quite frequently, details of their mechanisms have been less well documented: 1. It has been suggested that active transport systems for benzoate (Thayer and Wheelis 1982) and for mandelate (Higgins and Mandelstam 1972) are involved. In Rhizobium leguminosarum, 4-hydroxybenzoate hydroxylase activity is required for the uptake of 4-hydroxybenzoate (Wong et al. 1994), while in Pseudomonas putida a gene cluster pcaRKF is involved not only in the transport of 4-hydroxybenzoate into the cells, but also their chemotactic response to the substrate and its degradation by ring hydroxylation (Harwood et al. 1994; Nichols and Harwood 1997). The situation for phenylacetate transport into Pseudomonas putida U is apparently different (Schleissner et al. 1994), and this is consistent with the fact that the pathway for degradation of phenylacetate by this strain is distinct from that used for 4-hydroxyphenylacetate. 2. A detailed investigation by Groenewegen et al. (1990) has examined the uptake of 4-chlorobenzoate by a coryneform bacterium that degraded this compound. The uptake was inducible and occurred in cells grown with 4-chlorobenzoate but not with glucose. A proton motive force (∆p)–driven mechanism was almost certainly involved, and uptake could not take place under anaerobic conditions unless an electron acceptor such as nitrate was present. 3. The transport of toluene-4-sulfonate into Comamonas testosteroni has been examined (Locher et al. 1993), and rapid uptake required growth of the cells with toluene-4-sulfonate or 4-methylbenzoate. From the results of experiments with various inhibitors, it was concluded that a toluenesulfonate anion/proton symport system operates rather than transport driven by a difference in electrical potential (∆ψ), and uptake could not take place under anaerobic conditions unless an electron acceptor such as nitrate was present. 4. The uptake of benzoate was examined in two strains of Alcaligenes denitrificans. The transport system was inducible, carrier-mediated, energy-dependent, and involved a proton symport system. In contrast, the uptake of 2,4-dichlorobenzoate by one of the strains was constitutive, displayed no saturation kinetics, and appeared to occur by passive diffusion (Miguez et al. 1995). The uptake of 2,4-dichlorophenoxyacetate has been studied in Ralstonia eutropha JMP134(pJP4) in which the degradative genes are plasmid-borne.

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Uptake was inducible by the substrate, did not occur with fructose-grown cells, and was inhibited by cyanide that prevents development of a protomotive force, and by the protophore carbonylcyanide-3-chlorophenylhydrazone (Leveau et al. 1998). The protein involved was encoded by an open reading frame on the plasmid designated tfdK. Sphingobium (Sphingomonas) herbicidovorans strain MH is able to grow at the expense of both enantiomers of 2-(2,4-dichlorophenoxy)propionate with a preference for the S-enantiomer, and the uptake of both enantiomers, and of 2,4-dichlorophenoxyacetate is inducible. Although the ATPase inhibitor N,N′-dicyclohexylcarbodiimide (DCCD) had only slight effect on intracellular levels of ATP, uptake was inhibited by the protophore carbonylcyanide-3-chlorophenylhydrazone and by nigericin that dissipates ∆pH in the presence of high concentration of K+, but not by valinomycin. It was suggested that uptake is driven by a protomotive force (∆p), and that ∆pH rather than ∆ψ is the determinant of uptake (Zipper et al. 1998). In Oxalobacter formigenes, oxalate and its decarboxylation product formate form a oneto-one antiport system, which involves the consumption of an internal proton during decarboxylation, and serves as a proton pump to generate ATP by decarboxylative phosphorylation (Anantharam et al. 1989). The transport of EDTA into a bacterial strain capable of its degradation has been examined (Witschel et al. 1997). Inhibition was observed with DCCD (ATPase inhibitor), nigericin (dissipates ∆pH), but not valinomycin (dissipates ∆ψ), and was dependent on the stability constant of metal–EDTA complexes. A graphic example of the significance of effective transport is provided by an aerobic reductase from Xenophilus azovorans. This was expressed in Escherichia coli, and was able to reduce a range of important sulfonated colorants, even though whole cells were unable to do so (Blümel et al. 2002).

These results illustrate both the potential complexity and the significance of transport systems in bacteria. This has hitherto been a rather neglected aspect of the degradation of xenobiotics, and extension to other organisms and to a wider range of xenobiotics is clearly merited.

PRE-EXPOSURE: PRISTINE AND CONTAMINATED ENVIRONMENTS Although experimental aspects of the elective enrichment procedure are discussed in Chapter 5, the question of its existence and significance in natural populations already exposed to xenobiotics is addressed here. It is important to distinguish between induction (or derepression) of catabolic enzymes, and selection for a specific phenotype. The former is a relatively rapid response, so that exposure of samples from uncontaminated areas to xenobiotics for a period of weeks or months would be expected to result in the selection of organisms with degradative potential rather than merely be the result of low rates of enzyme induction. Caution should, however, be exercised in establishing a correlation between exposure to xenobiotics and the existence of organisms with the relevant degradative capacity. The following diverse examples are given as illustration: 1. It has been shown (Kamagata et al. 1997) that bacteria isolated from a pristine site with no established contamination were capable of degrading 2,4-dichlorophenoxyacetate (2,4-D) and differed from those traditionally isolated from contaminated sites. The new isolates grew slowly and although one of them could be assigned to the genus Variovorax and carried the tfdA gene, the other five did not and had no 2,4-D-specific 2-ketoglutaratedependent dioxygenase activity. 2. These results are consistent with previous evidence (Fulthorpe et al. 1996) from a study of pristine soils that although populations existed that could degrade both 3-chlorobenzoate

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and 2,4-dichlorophenoxyacetate, isolation of the latter strains was generally unsuccessful by the methods that were used. These results should be viewed against the general comments on oligotrophs and bradytrophs (slow-growing organisms) that have already been discussed. 3. Soils from putatively pristine areas in Southwest Australia; South Africa; California, the United States; Central Chile, Saskatchewan Canada; and Russia were enriched with 3-chlorobenzoate and assayed for mineralization of the substrate (Fulthorpe et al. 1998). Genetic procedures were used to show that 91% of the genotypes were unique to the sites from which the organisms were isolated. These results suggest that the genotypes were endemic and were not the result of global dispersion of single genotype. 4. Comparison has been made of the degradability in organisms isolated from a contaminated and a pristine system in Canada. It was shown that the established genotypes for the degradation of 3-chlorobenzoate i.e., clc, cba, and fcb that encode enzymes for 3-chlorocatechol 2,3-dioxygenase (Frantz and Chakrabarty 1987), 3,4-(4,5)-dioxygenase (Nakatsu and Wyndham 1993), and the hydrolytic pathway for 4-chlorobenzoate (Chang et al. 1992) were present in strains from the contaminated site. However, these were absent in strains isolated from the pristine site. Based on substrate utilization patterns, these traits were distributed among phenotypically distinct groups (Peel and Wyndham 1999). 5. Degradation of contaminants may occur with bacteria that have been isolated from pristine environments without established exposure to the contaminants, and exhibit no dependence on substrate concentration. For example, organisms from a previously unexposed forest soil were able to degrade 2,4,6-trichlorophenol at concentrations up to 5000 ppm, and terminal restriction fragment length polymorphism analysis revealed that at concentrations up to 500 ppm, the bacterial community was unaltered (Sánchez et al. 2004). The classic experiments by Tattersfield (1928) clearly showed that the biocidal effect of naphthalene decreased during successive application, and that this was due to the biodegradation of naphthalene. Since then, many results have accumulated that support the view that pre-exposure increases the number of organisms capable of degrading a given xenobiotic, though fewer attempts have been made to quantify the number of organisms involved. There is convincing evidence that exposure to unusual substrates in laboratory experiments elicits the synthesis of genes for their degradation (Mortlock 1982). Increasing support for the view that exposure to xenobiotics increases the probability of mutations that are favorable to the degradation of these substrates has also been found (Hall 1990; Thomas et al. 1992). The following examples attempt to illustrate the spectrum of responses that have been observed: 1. Rates of mineralization of the more readily degraded PAHs such as naphthalene and phenanthrene were greater in samples from PAH-contaminated areas than in those from pristine sediments, although it is significant that even in the former, the rates for benz[a]anthracene and benzo[a]pyrene were extremely low (Herbes and Schwall 1978). Examination of the sequences of 16S rRNA and of naphthalene dioxygenase [Fe-S] protein genes (nahAc) of nine strains of bacteria capable of degrading naphthalene and isolated from the same site showed that whereas in seven strains the dioxygenase differed by as much as 7.9%, they had a single nahAc allele. All strains contained plasmids of different sizes that contained the gene for naphthalene degradation, and it was suggested that horizontal transfer of plasmids might play a role in the adaptation of microbial communities to xenobiotics (Herrick et al. 1997). 2. Experiments using marine sediment slurries have examined the effect of pre-exposure to various aromatic hydrocarbons on the rate of subsequent degradation of the same, or other hydrocarbons. The results clearly illustrated the complexity of the selection process: for example, whereas pre-exposure to benzene, naphthalene, anthracene, or phenanthrene

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enhanced the rate of mineralization of naphthalene, similar pre-exposure to naphthalene stimulated the degradation of phenanthrene but had no effect on that of anthracene (Bauer and Capone 1988). In experiments using soil samples from a pristine aquifer exposed in the laboratory to a range of compounds, the following widely diverse responses were observed (Aelion et al. 1987, 1989): a. The bacterial population was apparently already adapted to some of the compounds such as phenol, 4-chlorophenol and 1,2-dibromoethane at the start of the experiment, and these substrates were therefore rapidly degraded. b. No adaptation was found for chlorobenzene or 1,2,4-trichlorobenzene, and only slight mineralization was observed. c. A linear increase in the rate of degradation with increasing length of exposure was noted for some substrates such as aniline, which are generally regarded as readily degradable. d. True adaptation by selection of the appropriate organisms was observed only for 4-nitrophenol. Systematic studies on the degradation of 4-nitrophenol (Spain et al. 1984) showed that the rates of adaptation in a natural system were comparable to those observed in a laboratory test system and were associated with an increase in the number of degrading organisms by up to 1000-fold. It has been consistently observed that a wide range of agrochemicals applied successively to the same plots are increasingly readily degraded, presumably due to enrichment of the appropriate degradative microorganisms (Racke and Coats 1990). Examples include compounds as diverse as naphthalene (Gray and Thornton 1928), γ-hexachloro[aaaeee]cyclohexane (Wada et al. 1989), and triazines (Cook and Hütter 1981). The dehalogenation of polychlorinated or polybrominated biphenyls was more rapid in cultures using inocula prepared from sediments contaminated with the chlorinated or brominated biphenyl, respectively (Morris et al. 1992). The enrichment cultures from soils treated with the urea herbicide linuron for more than 10 years showed specific degradation of linuron and the related bromuron, though not the urea herbicides lacking the methoxy group. Samples from untreated soil did not display this behavior (El-Fantroussi 2000). The use of reverse transcription-PCR and denaturing gradient gel electrophoresis showed that a bacterial consortium was required for the complete degradation of linuron. Variovorax sp. strain WDL1, Delftia acidovorans WDL34, and Pseudomonas sp. strain WDL5 could be isolated directly from the enrichment, and Hyphomicrobium sulfonivorans WDL6 and Comamonas testosteroni strain WDL7 from subenrichments (Dejonghe et al. 2003). Only Variovorax sp. WDL1 was, however, able to use linuron as the sole source of carbon and energy. Long-term laboratory enrichments have been shown to yield cultures that were able to degrade initially recalcitrant compounds. Several examples have been given by Slater, who was a pioneer in this area (Slater and Lovatt 1984). An additional illustration is provided by an organism that was able to grow at the expense of 4,4′-dicarboxyazobenzene. This had been isolated during prolonged continuous cultivation (≈500 d) under nonsterile conditions using the more complex naphthalene analogs (Kulla 1981). Pure cultures of Xenophilus azovorans KF46 and Pigmentiphaga kullae were then isolated from the adapted culture (references in Blümel et al. 2002).

It therefore seems premature to draw general conclusions on the influence of pre-exposure on the biodegradability of structurally different substrates. These results are particularly relevant to bioremediation, and suggest that organisms originally isolated from sites either contaminated naturally (Fredrickson et al. 1991) or as a result of industrial

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activity (Grosser et al. 1991) may be particularly suitable. The discussion hitherto has been devoted to the issue of selection, but for the sake of completeness it should be noted that the alternative approach of deliberately adding an inducer has also been examined. Salicylate is an inducer of the enzymes for degradation of naphthalene. Its addition to soil has been shown to result in a modest increase in the number of organisms degrading naphthalene (Ogunseitan et al. 1991), and in the mineralization of some 4- and 5-ring PAHs (Chen and Aitken 1999). However, since naphthalene and phenanthrene are readily degradable compounds, the wider evaluation of this interesting idea has to be explored before its application to practical situations can be justified. Substrates whose structures are unrelated to the contaminant have been noted for PCB degradation: (a) growth on a range of flavones, flavanols, and flavanones (catechin, naringenin, and myricitin) (Donnelly et al. 1994) and (b) induction by terpenoids including carvone and limonene (Gilbert and Crowley 1997). This discussion may also be viewed against the background of studies on genetic transfer of catabolic activity toward a given xenobiotic. Considerable care should, however, be exercised in the interpretation even of data that seem supportive of this process, since selection and enrichment from a small population of organisms initially present must be excluded. One interesting investigation on the biodegradation of aniline revealed the existence of two genotypes differing in their tolerance to the substrate. The dominant organism that was originally present assimilated aniline at micromolar concentrations, although it was inhibited at higher concentrations. A mutant could, however, be isolated from a population of several hundred cells or by continuous culture, and this organism tolerated millimolar concentrations. Populations of the two organisms in a natural system were apparently regulated by the prevailing concentration of aniline (Wyndham 1986). Although great advances have been made using the technology of molecular biology to determine the components of natural populations without the limitations of specific enrichment, it is hardly possible to escape the conclusion that our understanding of the processes regulating the population dynamics of microorganisms in natural systems is still limited. Details of procedures for investigating natural populations in the context of bioremediation are addressed in Chapter 13.

RATES OF METABOLIC REACTION KINETIC ASPECTS The rates at which xenobiotics are degraded or transformed are of cardinal importance, since it is upon their quantification that a given compound can, in the final analysis be designated persistent or otherwise (Battersby 1990). It should be appreciated at the outset that even if acceptable rates of degradation are observed in laboratory experiments, the final assessment of persistence depends upon the demonstration that the compound is indeed degraded under natural conditions. Ultimately, this is of primary environmental significance. It has already been pointed out that persistence is determined not only by rates of biotic and abiotic degradation, but also by the accessibility of the substrate that may be associated with organic or inorganic components in the water mass, or in the soil or sediment phases. There are a number of additional issues that should be addressed in the discussions of rates: 1. Even if rate constants can be measured in laboratory experiments, these must be normalized to the number of microbial cells. This may pose only minor problems with an axenic culture in the laboratory, and has been consistently carried out in investigations where well-defined kinetics prevailed (Allard et al. 1987). However, this becomes a major problem in natural situations: how many of the organisms are metabolically active in accomplishing the given reaction? Use of specific DNA probes has been used for detection of genes coding for heavy-metal resistance (Diels and Mergeay 1990), and for the detection of pathogens (Samadapour et al. 1990). This procedure has, however, been less extensively applied to organisms of catabolic significance (Sayler et al. 1985; Holben et al. 1988), although the values obtained for 2,4-dichlorophenoxyacetic-degrading

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populations agreed well with those using conventional most-probable-number methods (Holben et al. 1992). 2. Although relevant models have been assembled (Simkins and Alexander 1984) and their application evaluated (Simkins et al. 1986), microbial reactions may not follow welldefined kinetics. They may, for example, exhibit multiphase kinetics that has been illustrated during the transformation of methyl parathion by a Flavobacterium sp. (Lewis et al. 1985). System I was a high-affinity, low-capacity system, whereas system II was the opposite. In experiments with mixed microbial cultures on the degradation of phenol, 4-chlorophenol, 4-methylphenol, acetone, and methanol, multiphase kinetics were encountered. Failure to take this into account would have resulted in serious errors in evaluating the rates of degradation (Hwang et al. 1989). The degradation of 1,2,4-trichlorobenzene by Burkholderia sp. strain PS14 at low substrate concentrations showed a first-order relation between the specific rate of transformation and the substrate concentration. However, at higher concentration, this was replaced by a second-order one (Rapp 2001). The results of these studies illustrate only one of the several factors that may invalidate predictions that fail to take into account multiphase kinetics. 3. Attempts have been made to apply the structure-activity concept (Hansch and Leo 1995) to environmental problems, and this has been successfully applied to the rates of hydrolysis of carbamate pesticides (Wolfe et al. 1978), and of esters of chlorinated carboxylic acids (Paris et al. 1984). This has been extended to correlating rates of biotransformation with the structure of the substrates and has been illustrated with a number of single-stage reactions. Clearly, this approach can be refined with the increased understanding of the structure and function of the relevant degradative enzymes. Some examples illustrate the application of this procedure: a. Rates of bacterial hydroxylation of substituted phenols to catechols by Pseudomonas putida correlated well with the van der Waals radii of the substituents (Paris et al. 1982). This was also demonstrated for the biotransformation of anilines to catechols both by this strain and by a natural population of bacteria (Paris and Wolfe 1987). b. Rates of hydrolysis of substituted aromatic amides by the bacterial population of pond water correlated well with the infrared C=O stretching frequencies of the substrates (Steen and Collette 1989). c. Rates of anaerobic dechlorination of aromatic hydrocarbons (Peijnenburg et al. 1992), and of the hydrolysis of aromatic nitriles under anaerobic conditions (Peijnenburg et al. 1993) have been correlation with a number of parameters including Hammett σ-constants, inductive parameters, and evaluations of the soil/water partitioning. For a number of reasons, there are some important limitations to the extension of this principle. Biodegradation—as opposed to biotransformation—of complex molecules necessarily involves a number of sequential reactions each of whose rates may be determined by complex regulatory mechanisms. For novel compounds containing structural entities that have not been previously investigated, the level of prediction is necessarily limited by lack of the relevant data. Too Olympian a view of the problem of rates should not, however, be adopted. An overly critical attitude should not be allowed to pervade the discussions—provided that the limitations of the procedures that are used are clearly appreciated and set forth. In view of the great practical importance of quantitative estimates of persistence to microbial attack, any procedure—even if it provides merely orders of magnitude—should not be neglected.

METABOLIC ASPECTS: NUTRIENTS In natural ecosystems, microbial growth and metabolism may be limited by the concentrations of inorganic nutrients such as nitrogen, phosphorus, or even iron. Systematic investigation of these

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limitations on the biodegradation of xenobiotics has seldom been carried out except when the substrates contain nitrogen or phosphorus in a form that is used for growth of the cells. In growth media containing nitrate, unusual metabolites containing nitro groups have been found. These have been discussed in Chapter 2. The role of Cu, Ni, Mo, and W are discussed in the context of metalloenzymes in Chapter 3, Part 5. Only two examples of the importance of nutrient limitation in determining the realization of biodegradation are noted here: 1. Expression of lignin peroxidases in Phanerochaete chrysosporium is induced by nitrogenlimitation, and by the concentration of Mn (II) in the medium (Perez and Jeffries 1990). 2. Under conditions of selenium starvation, Clostridium purinilyticum degrades uric acid by an unusual pathway involving cleavage of the iminazole ring to produce 5,6-diaminouracil, which is then degraded to formate, acetate, glycine, and CO2 (Dürre and Andreesen 1982). In experiments using field samples containing natural assemblies of microorganisms, at least two effects of supplementation with nitrogen or phosphorus have been encountered: 1. A decreased lag phase was observed before transformation of 4-methylphenol (Lewis et al. 1986). 2. The size of the bacterial population increased, although there was no effect on the secondorder rates of transformation of a number of compounds including phenol, and the agrochemicals propyl 2,4-dichlorophenoxyacetate, methyl parathion and methoxychlor (Paris and Rogers 1986). Limitation in the concentrations of these inorganic nutrients does not therefore appear to have had a dramatic effect on the persistence of the relatively few compounds that have been examined systematically. However, it has been suggested that competition between organisms for inorganic phosphorus may account for differences in biodegradability when several carbon sources are present (Steffensen and Alexander 1995). Consistent with the preceding comments on the metabolism of xenobiotics in the presence of additional carbon substrates, the result of deliberate addition of organic carbon may be quite complex and will not be addressed in detail. Two examples on rates of mineralization are given as illustration in which addition of glucose apparently elicited two different responses. It should, however, be emphasized that since the concentration of readily degradable substrates in natural aquatic systems will generally be extremely low, the environmental relevance of such observations will inevitably be restricted: 1. The rate of mineralization of phenol by the flora of natural lake water decreased (Rubin and Alexander 1983). 2. The rate of mineralization of 4-nitrophenol was enhanced in lake water inoculated with a Corynebacterium sp. when rates of mineralization were low (Zaidi et al. 1988).

REGULATION AND TOXIC METABOLITES REGULATION Whereas the following discussion is directed primarily to the role of metabolites, environmental factors may also be important. Some examples including the role of oxygen concentration have already been given as illustration. Classic studies were devoted to the regulation of the enzymes for conversion of catechol and 3,4-dihydroxybenzoate (protocatechuate) to 3-ketoadipate by Pseudomonas putida (Ornston 1966),

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the 3-ketoadipate pathway in Moraxella calcoacetica (Cánovas and Stanier 1967), and the mandelate pathway in Pseudomonas aeruginosa (Rosenberg 1971). In these organisms, catechol is degraded by intradiol fission catalyzed by a 1,2-dioxygenase, and the induction patterns are briefly summarized. Synthesis of the enzymes for the primary oxidations is induced by the growth substrates, whereas synthesis of those for later transformations is induced by their products—either by cis,cismuconate, or by 3-ketoadipate. However, for Alcaligenes eutrophus that metabolizes phenol and 4-methylphenol by extradiol fission catalyzed by catechol 2,3-dioxygenase, all six enzymes that result in the production of 4-hydroxy-2-ketovalerate are inducible by the initial growth substrates (Hughes and Bayly 1983). The induction patterns may, however, be quite complex and depend strongly on the growth substrate. For example, Pseudomonas paucimobilis (strain Q1) that was isolated by enrichment with biphenyl is capable of utilizing biphenyl, xylene and toluene, salicylate, and benzoate (Furukawa et al. 1983). All the enzymes for degradation of the first four of these that are degraded by 2,3-dioxygenase ring fission were induced by growth on these substrates, whereas growth with benzoate induced only the 1,2-dioxygenase rather than the 2,3-dioxygenase. It is instructive to note that whereas the enzymes of the 3-ketoadipate pathway are generally inducible, this is not always the case: those in slow-growing species of Bradyrhizobium spp. are constitutive (Parke and Ornston 1986). Except for aromatic hydrocarbons and related compounds, somewhat fewer studies have been directed to the genetics and the regulation of the enzymes for the degradation of xenobiotics. The following are illustrative examples that have received extensive study: a. Pseudomonas putida strain F1 that degrades toluene with the methyl group intact (Finette et al. 1984). b. For the nah operon in Pseudomonas putida strain G1064 involved in the degradation of naphthalene to salicylate (Eaton and Chapman 1992 ), the enzymes are generally induced by growth with salicylate (Austen and Dunn 1980). c. The initial steps in the degradation of biphenyls by Pseudomonas sp. strain LB400 (Mondello 1989), by Pseudomonas pseudoalcaligenes strain KF707 (Taira et al. 1992), and by Rhodococcus globerulus (Asturias and Timmis 1993). This is discussed in greater detail in Chapter 9, Part 2. Further details of the pathways for the degradation of PAHs are given in Chapter 8, Part 1, and in reviews that are cited there. It seems that most of the degradative enzymes are inducible, and this is consistent with the fact that most strains have been isolated after specific enrichment with the xenobiotic. Synthesis of catechol 1,2-dioxygenase in the yeast Trichosporon cutaneum is, however, partially constitutive (Shoda and Udaka 1980). For biotransformation, however, there are sporadic examples of the constitutive synthesis of enzymes. For example, the system carrying out the O-methylation of halogenated phenolic compounds was apparently constitutive (Neilson et al. 1988), and is consistent with the isolation of the strains by enrichment with C1 compounds structurally unrelated to the halogenated substrates. The O-methylation reaction may function primarily as a detoxification system, so that in this case constitutive synthesis of the enzyme would clearly be advantageous to the survival of the cells. Since it is only seldom that organisms are exposed to a single substrate, attention has been directed to the metabolism of mixtures of benzoate and aliphatic carboxylic acids that may be the products of degradation. In Ralstonia eutropha (Alcaligenes eutrophus), diauxic growth was exhibited when both benzoate and acetate were supplied as substrates, though not when benzoate and succinate were used (Ampe et al. 1997). When both of these were supplied, the growth rate was increased, and under these circumstances there was putatively a more optimal distribution of metabolites among the various anabolic and catabolic pathways. In contrast, the simultaneous presence of acetate and benzoate would result in the production of energy that would exceed the requirements of the cell. However, during growth with both acetate and phenol, synthesis of both phenol hydroxylase and the extradiol dioxygenase involved in the metabolism of catechol decreased, with the result that benzoate blocked diauxic growth with phenol (Ampe et al. 1998).

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An interesting and ecologically relevant observation has been made on the induction of the metabolism of PCBs by Arthrobacter sp. strain B1B (Gilbert and Crowley 1997). Cells were grown in a mineral medium with fructose and carvone (50 mg/L). Effective degradation of a number of congeners of Arochlor 1242 was induced by carvone that could not, however, be used as a growth substrate and was toxic at high concentrations (> 500 mg/L). Other structurally related compounds including limonene, p-cymene, and isoprene were also effective, and such results may be relevant to bioremediation programs for PCBs. It has also been shown by protein sequencing in Comamonas testosteroni that testosterone induces both the synthesis of enzymes that degrade steroids and those required for the degradation of aromatic compounds (Möbus et al. 1997). The induction of the monooxygenases for the degradation of trichloroethene by aromatic substrates and vice versa is discussed in Chapter 8, Part 1.

TOXIC OR INHIBITORY METABOLITES There are several examples in which metabolites that toxify the organism responsible for their synthesis are produced. The classic example is fluoroacetate (Peters 1952), which enters the TCA cycle and is thereby converted into fluorocitrate. This effectively inhibits aconitase—the enzyme involved in the next metabolic step—so that cell metabolism itself is inhibited with the resulting death of the cell. Walsh (1982) has extensively reinvestigated the problem and revealed both the complexity of the mechanism of inhibition and the stereospecificity of the formation of fluorocitrate from fluoroacetate (p. 239). It should be noted, however, that bacteria able to degrade fluoroacetate to fluoride exist so that some organisms have developed the capability for overcoming this toxicity (Meyer et al. 1990). The significance of toxic metabolites is important in diverse metabolic situations: (a) when a pathway results in the synthesis of a toxic or inhibitory metabolite, and (b) when pathways for the metabolism of two (or more) analogous substrates supplied simultaneously are incompatible due to the production of a toxic metabolite by one of the substrates. A number of examples are provided to illustrate these possibilities that have achieved considerable attention in the context of the biodegradation of chlorinated aromatic compounds (further discussion is given in Chapter 9, Part 1): 1. Biotransformation of chlorobenzene by Pseudomonas putida grown with toluene or benzene resulted in the formation of 3-chlorocatechol. This inhibited further metabolism by catechol 2,3-dioxygenase, so that its presence resulted in the formation of catechols even from benzene and toluene (Gibson et al. 1968; Klecka and Gibson 1981; Bartels et al. 1984). The same situation has emerged in the degradation of 3-chlorobenzoate: the inhibition that would result form the inhibitory effect of 3-chlorocatechol on catechol 2,3dioxygenase is lifted by the synthesis of catechol 1,2-dioxygenase (Reinecke et al. 1982). This is indeed a general strategy, and inactivation of extradiol dioxygenases—reversible or otherwise—by other substituted catechols has also been reported: a. 2,3-Dihydroxybiphenyl 1,2-dioxygenase from a biphenyl-degrading strain of Pseudomonas sp. by 4-phenyl catechol that is not a substrate for the dioxygenase (Lloyd-Jones et al. 1995) b. Catechol 2,3-dioxygenase by 4-ethyl catechol that is a metabolite of 4-ethylbenzoate (Ramos et al. 1987) Although this inactivation may present a limitation in the degradative potential of the relevant strains, Pseudomonas putida strain GJ31 degrades chlorobenzene via 3-chlorocatechol and extradiol fission (Kaschabek et al. 1998). This is accomplished by a chlorocatechol 2,3-dioxygenase that hydrolyses the initially formed cis,cis-hydroxy muconic acid chloride to 2-hydroxymuconate. Thereby the irreversible reaction of the acid chloride with nucleophiles or the formation of pyr-2-one-6-carboxylate as a terminal metabolite are avoided. The catechol 2,3-dioxygenase from this strain encoded by cbzE is plasmid-borne, and is capable of metabolizing both 3-chlorocatechol and 3-methylcatechol. It belongs to the

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2.C subfamily of type 1 extradiol dioxygenases (Mars et al. 1999). The alternative extradiol fission of 3-chlorocatechol may take place between the 1 and 6 positions (distal fission), and this has been shown for the 2,3-dihydroxybiphenyl 1,2-dioxygenase from the naphthalene sulfonate degrading Sphingomonas sp. strain BN6 (Riegert et al. 1998). The converse situation occurs during the degradation of 4-chlorobenzoate in which the synthesis of catechol 1,2-dioxygenase circumvents the production of chloroacetaldehyde by the action of catechol 2,3-dioxygenase (Reinecke et al. 1982). It has been shown that the synthesis of protoanemonin may take place during the degradation of 4-chlorobenzoate, which is formed by ring fission of 4-chlorobiphenyl (Blasco et al. 1995), and that this may inhibit the growth of PCB-degrading organisms in the soil (Blasco et al. 1997). Although protoanemonin is produced from chloro-cis-cis-muconates, its formation may be circumvented by the action of chloromuconate cycloisomerases to form dienlactones, and by conversion by dienelactone hydrolase into cis-acetylacrylate (Brückmann et al. 1998) (Figure 4.11). 2. Ring fission during the degradation of chloroaromatic compounds is generally catalyzed by catechol-1,2-dioxygenases, whereas for the corresponding methyl compound a catechol-2,3dioxygenase is involved. These pathways are generally incompatible due to the inactivation of catechol-2,3-oxygenase by 3-chlorocatechol. Simultaneous degradation of chloro- and methylsubstituted aromatic compounds has, however, been shown to occur in some strains (Taeger 1988; Pettigrew et al. 1991), while a strain of Ralstonia sp. strain JS705 carries genes for both chlorocatechol degradation and toluene dioxygenation (Van der Meer et al. 1998). Although Comamonas testosteroni strain JH5 degraded 4-chlorophenol and 2-methylphenol simultaneously, 4-chlorophenol and 4-methylphenol were degraded sequentially. This involved a catechol-2,3-dioxygenase for both 4-methylcatechol and 4-chlorocatechol that was induced by growth with 4-chlorophenol or 4-methylphenol (Hollender et al. 1994). Phanerochaete chrysosporium is able to degrade simultaneously chlorobenzene and toluene (Yadav et al. 1995). This phenomenon is highly relevant to the biological treatment of industrial effluents, since most of these consist of complex mixtures of substrates, and to bioremediation. 3. Total degradation of PCBs necessitates degradation of the chlorobenzoates produced by dioxygenation, dehydrogenation, and ring fission. Metabolism of chlorobenzoate may produce chlorocatechols and thence muconic acids. It has been shown that in Pseudomonas testosteroni strain B-356, the metabolites from 3-chlorobenzoate strongly inhibited the activity of 2,3-dihydroxybiphenyl 1,2-dioxygenase, and therefore the degradation of the original PCB substrates (Sondossi et al. 1992). This inhibition is reminiscent of the inhibition of catechol 2,3-dioxygenase by 3-chlorocatechol, which has already been noted.

O CO2H CO2 H

O OH

O OH

CO2H

CH 3 CO2 H

Cl

Cl

O CO2 H

Cl O

O CO2 H

FIGURE 4.11

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CO2 H

Biodegradation of 4-chlorobenzoate.

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4. Simultaneous metabolism of even closely related substrates may be restricted by the synthesis of inhibitory metabolites. For example, cells of an Acinetobacter sp. grown with 4-chlorobenzoate could dehalogenate 3,4-dichlorobenzoate even though the organism cannot use this as the sole growth substrate, and the dichlorobenzoate inhibited both its own metabolism and that of the growth substrate 4-chlorobenzoate (Adriaens and Focht 1991a,b). 5. A strain of Pseudomonas sp. grew with a wide range of aromatic compounds including phenol, benzoate, benzene, toluene, naphthalene, and chlorobenzene (Haigler et al. 1992), and mixtures of these substrates were degraded in continuous cultures without evidence for accumulation of metabolites. In this case, degradation depended on the presence of a nonspecific toluene dioxygenase, and of induction of enzymes for the intradiol (ortho) fission, extradiol (meta) fission, and the “modified intradiol fission (ortho) pathway” for the degradation of the catechols. The results implied the existence of a metabolic situation in addition to that normally encompassed by the terms biodegradation and biotransformation—enzymes induced by the presence of one substrate facilitated the degradation of another substrate that is not normally used as the sole source of carbon and energy. These results have obvious implications for implementing bioremediation programs. 6. There may be several reasons why an analog substrate cannot be metabolized by an organism. This is well illustrated by 4-ethylbenzoate that could not be used by strains that degraded 3- and 4-methylbenzoate (Ramos et al. 1987). There were two reasons: a. The metabolite 4-ethylcatechol inactivated catechol 2,3-dioxygenase that is obligatory for its degradation. b. 4-Ethylcatechol, in contrast to 3- and 4-methylcatechol, does not activate the xylS protein whose gene is the positive regulator of the promoter of the TOL plasmid extradiol fission pathway. These limitations could, however, be overcome by the construction of mutant strains. 7. The degradation of trichloroethene by methylotrophic bacteria involves the epoxide as intermediate (Little et al. 1988). Further transformation of this may produce CO that can toxify the bacterium, both by competition for reductant and by enzyme inhibition (Henry and Grbic-Galic 1991). The inhibitory effect of CO may, however, be effectively overcome by adding a reductant such as formate. 8. The growth of a strain of Alcaligenes sp. that could degrade 2-aminobenzenesulfonate was inhibited by equimolar concentrations of 3-methylbenzoate, although an exconjugant after insertion of the TOL catabolic genes was able to carry out sequential degradation of both substrates (Jahnke et al. 1993).

CATABOLIC PLASMIDS Plasmids may be defined as fragments of DNA that replicate outside the bacterial chromosome and they are important in a number of different contexts: 1. As carriers of antibiotic resistance (Chapter 3, Part 4): the emergence of antibiotic-resistant strains has had serious repercussions in the application of antibiotic therapy, and has seriously increased the danger of nosicomial infections. 2. The presence of unusual carbohydrate fermentation patterns (particularly for lactose), and the ability to use citrate among Enterobacteriaceae has hindered, and sometimes jeopardized, the identification of pathogenic strains including Salmonella typhi. 3. Resistance to heavy metals including Hg may be mediated by plasmid-borne genes. Further discussion of resistance to metal cations and metalloid anions is given in Chapter 3, Part 4.

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4. Genes coding for the catabolism of a large number of diverse xenobiotics are often carried on plasmids. This is highly germane to the present discussion, and some representative examples are given in Table 4.2. Degradative plasmids for substrates including PAHs and chlorophenolic compounds are widely distributed among strains of Sphingomonas (Basta et al. 2004). A large number of them were megaplasmids, although some degradative genes were also chromosomal. Plasmid transfer may occur within the genus Sphingomonas and closely related genera, but apparently infrequently to other genera. In addition, many bacteria harbor plasmids with no hitherto established function.

TABLE 4.2 Examples of the Range of Plasmids Carrying Genes Coding Enzymes for the Biodegradation or Biotransformation of Xenobiotics Reference Hydrocarbons and related compounds Octane Benzene Toluene and xylene Naphthalene Naphthalene, phenanthrene, and anthracene Dibenzothiophene

Chakrabarty et al. (1973) Tan and Mason (1990) Williams and Worsey (1976) Connors and Barnsley (1982) Sanseverino et al. (1993) Monticello et al. (1985)

Terpenoids Camphor Citronellol, geraniol Linalool

Rheinwald et al. (1973) Vandenbergh et al. (1983) Vandenbergh et al. (1986)

Halogenated compounds Bromoxynil Chloroalkanoates Chlorophenoxyacetates Dichloromethane Chlorobenzenes Chlorobenzoates Chlorobiphenyls Chloridazon γ-Hexachlorocyclohexane

Stalker and McBride (1987) Kawasaki et al. (1981); Hardman et al. (1986) Don and Pemberton (1985); Chaudhry and Huang (1988) Gälli and Leisinger (1988) Van der Meer et al. (1991) Chatterjee et al. (1981) Furukawa and Chakrabarty (1982); Shields et al. (1985) Kreiss et al. (1981) Miyazaki et al. (2006)

Other structural groups 6-Aminohexanoate cyclic dimer Parathion Nicotine Cinnamic acid Ferulic acid S-ethyl-N,N-dipropylthiocarbamate Aniline 2-Aminobenzenesulfonate 1,1′-Dimethyl-4,4′-bipyridinium dichloride Arylsulfonates

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Negoro et al. (1980); Kanagawa et al. (1989) Serdar et al. (1982); Mulbry et al. (1986) Brandsch et al. (1982) Andreoni and Bestetti (1986) Andreoni and Bestetti (1986) Tam et al. (1987) Anson and Mackinnon (1984); McClure and Venables (1987); Karns and Eaton (1997) Jahnke et al. (1990) Salleh and Pemberton (1993) Junker and Cook (1997)

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Genes for specific substrates may be borne either on plasmids or chromosomally: 1. In the degradation of toluene, the plasmid-borne genes encode enzymes for the oxidation to benzoate, whereas those that are chromosomally borne encode ring 2,3-dioxygenation. 2. The degradation of trichloroethene by Alcaligenes eutrophus JMP134 is carried out by either a chromosomal phenol-dependent pathway, or by the plasmid-borne 2,4-dichlorophenoxyacetate pathway (Harker and Kim 1990). 3. The degradation of 2,4-dichlorophenoxyacetate is initiated by a 2-ketoglutarate-dependent dioxygenase that results in the production of 2,4-dichlorophenol. The genes are generally plasmid-borne, although chromosomally borne genes have been identified in Burkholderia sp. strain RASC (Suwa et al. 1996). The transfer of both the plasmid-borne and chromosomally borne tfdA gene may contribute to the dissemination of capability to degrade 2,4-dichlorophenoxyacetate (Matheson et al. 1996; McGowan et al. 1998). Probably the greatest interest in catabolic plasmids stems from the possibility of constructing strains with increased metabolic potential toward xenobiotics, and from the potential application of such strains to waste-treatment systems. Considerable effort has been devoted to aromatic chlorinated compounds, and conspicuous success has been achieved in overcoming problems with, for example, the synthesis of toxic metabolites such as 3-chlorocatechol during the degradation of chlorobenzenes by natural strains. There are several important issues that should, however, be addressed before considering the application of such artificially constructed strains to biological treatment systems or bioremediation programs: 1. Competition with endemic strains, which may eventually outnumber and eliminate the introduced strains. 2. Genetic instability if selection pressure is removed as a result of fluctuations in the loading of the xenobiotic. 3. Concern over the discharge of such strains into the environment. Attempts have therefore been made to incorporate safeguards so that the strains are unlikely to survive in competition with natural strains in the ecosystem. Plasmid transmission and the stability of plasmids in natural ecosystems have received considerable attention, but caution should be exercised in drawing general conclusions on the basis of the sometimes fragmentary evidence from laboratory experiments. Some important principles are illustrated by the following: 1. A study with a strain of plasmid-borne antibiotic-resistant Escherichia coli indicated that the strain did not transmit these plasmids to indigenous strains after introduction into the terrestrial environment (Devanas et al. 1986). 2. It has been shown (Smith et al. 1978) that in enteric bacteria carrying thermosensitive plasmids coding for the utilization of citrate and for resistance to antibiotics, rates of transmission were negligible at 37°C but appreciable at 23°C—a temperature more closely approaching that which prevails in natural ecosystems. 3. There is evidence that even in the absence of selective pressure exerted by the presence of a xenobiotic, bacterial populations may retain a small number of organisms carrying the relevant degradative plasmids. For example, strains of Pseudomonas putida in which the degradation of toluene is mediated by genes on the nonconjugative TOL plasmid maintain a small population of cells carrying the plasmid even in the absence of toluene (Keshavarz et al. 1985). It has also been shown (Duetz et al. 1994) that cultures of Pseudomonas putida grown with growth-limiting concentrations of succinate express TOL catabolic genes that are responsible for degradation by both the upper and lower pathways. This was observed

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in response to o-xylene that is not metabolized, although not during nonlimiting growth with succinate. 4. Strains of indigenous groundwater bacteria and a strain of Pseudomonas putida carrying a TOL plasmid were introduced into microcosms prepared from a putatively pristine aquifer. DNA-specific probes were used to monitor the numbers of organisms carrying the genotypes and it was found that the stability of genotypes for the degradation of toluene was maintained in the absence of selective pressure over the 8-week period of the experiment (Jain et al. 1987). 5. A strain of Alcaligenes sp. carrying a plasmid bearing the genes for the degradation of 3-chlorobenzoate was introduced into a freshwater microcosm system, and a specific DNA probe was used to enumerate the organisms bearing this gene (Fulthorpe and Wyndham 1989). A number of generally important results were obtained: a. The presence of 3-chlorobenzoate was needed to maintain the catabolic genotype. b. The number of probe-positive organism often greatly outnumbered that of the original organism determined by plate counts; c. The nature of these “additional” probe-positive organisms was not established. 6. Detailed analysis of bacteria isolated from the sediment of a coal-tar-contaminated site demonstrated horizontal transfer of genes involved in naphthalene degradation, and that identical alleles NahAc were shared among seven taxonomically diverse hosts (Herrick et al. 1997). Analysis of restriction fragment length polymorphism (RFPL) revealed the existence of two catabolic plasmids in 12 of the isolates, and that these were closely related both to each other and to the plasmid pDGT1 of Pseudomonas putida strain NCIB 9816-4. The plasmids were stable in this environment, and it was therefore suggested that these plasmids were involved in the development of the naphthalene-degrading community as a result of the selective pressure exerted by the coal tar contamination (Stuart-Keil et al. 1998). Clearly, therefore further investigation is required before generalizations can be made on the cardinal issues of the stability of plasmids in natural ecosystems, the extent to which these plasmids are transmissible, and the stability of the genotypes in the absence of selective pressure. At least some of the apparently conflicting views may be attributed to the different organisms that have been used including their nutritional demands,compounded by the widely varying environments in which their stability have been examined (Sobecky et al. 1992). Currently, the greatest volume of research is being devoted to primarily genetic aspects of these problems, particularly in biotechnology.

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Williams WA, RJ May (1997) Low-temperature microbial aerobic degradation of polychlorinated biphenyls in sediment. Environ Sci Technol 31: 3491–3496. Williams PA, MJ Worsey (1976) Ubiquity of plasmids in coding for toluene and xylene metabolism in soil bacteria: evidence for the existence of new TOL plasmids. J Bacteriol 125: 818–828. Witschel M, S Nagel, T Egli (1997) Identification and characterization of the two-enzyme system catalyzing the oxidation of EDTA in the EDTA-degrading bacterial strain DSM 9103. J Bacteriol 179: 6937–6943. Wolfe NL, RG Zepp, DF Paris (1978) Use of structure–reactivity relationships to estimate hydrolytic persistence of carbamate pesticides. Water Res 12: 561–563. Wolin MJ (1982) Hydrogen transfer in microbial communities. In Microbial interactions and communities (Eds AT Bull, JH Slater), Vol. 1, pp. 323–356. Academic Press, New York. Wong CM, MJ Dilworth, AR Glenn (1994) Cloning and sequencing show that 4-hydroxybenzoate hydroxylase Poba is required for uptake of 4-hydroxybenzoate in Rhizobium leguminosarum. Microbiology (UK) 140: 2775–2786. Wu Q, DL Bedard, J Wiegel (1997a) Effect of incubation temperature on the route of microbial reductive dechlorination of 2,3,4,6-tetrachlorobiphenyl in polychlorinated biphenyl (PCB)-contaminated and PCB-free freshwater sediments. Appl Environ Microbiol 63: 2836–2843. Wu Q, DL Bedard, J Wiegel (1997b) Temperature determines the pattern of anaerobic microbial dechlorination of Arochlor 1260 primed by 2,3,4,6-tetrachlorobiphenyl in Woods Pond sediment. Appl Environ Microbiol 63: 4818–4825. Wyndham RC (1986) Evolved aniline catabolism in Acinetobacter calcoaceticus during continuous culture of river water. Appl Environ Microbiol 51: 781–789. Yadav JS, RE Wallace, CA Reddy (1995) Mineralization of mono- and dichlorobenzenes and simultaneous degradation of chloro- and methyl-substituted benzenes by the white-rot fungus Phanerochaete chrysosporium Appl Environ Microbiol 61: 677–690. Yu Z, GR Stewart, W Mohn (2000) Apparent contradiction: psychrotolerant bacteria from hydrocarboncontaminated arctic tundra soils that degrade diterpenoids synthesized by trees. Appl Environ Microbiol 66: 5148–5154. Zaidi BR, Y Murakami, M Alexander (1988) Factors limiting success of inoculation to enhance biodegradation of low concentrations of organic chemicals. Environ Sci Technol 22: 1419–1425. Zhang Y, RM Miller (1992) Enhanced octadecane dispersion and biodegradation by a Pseudomonas rhamnolipid surfactant (biosurfactant). Appl Environ Microbiol 58: 3276–3282. Zhang Y, WJ Maier, RM Miller (1997) Effect of rhamnolipids on the dissolution, bioavailability and biodegradation of phenanthrene. Environ Sci Technol 31: 2211–2217. Zhao H, D Yang, CR Woese, MP Bryant (1990) Assignment of Clostridium bryantii to Syntrophospora bryantii gen. nov., comb. nov. on the basis of a 16S rRNA sequence analysis of its crotonate-grown pure culture. Int J Syst Bacteriol 40: 40–44. Zipper C, M Bunk, AJB Zehnder, H-PE Kohler (1998) Enantioselective uptake and degradation of the chiral herbicide dichloroprop [(RS)-2-(2,4-dichlorophenoxy)propionic acid] by Sphingomonas herbicidovorans MH. J Bacteriol 180: 3368–3374. Zylstra GJ, LP Wackett, DT Gibson (1989) Trichloroethylene degradation by Escherichia coli containing the cloned Pseudomonas putida F1 toluene dioxygenase genes. Appl Environ Microbiol 55: 3162–3166.

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INTRODUCTION A general overview of the processes that determine the fate and persistence of xenobiotics in the environment has been presented in Chapters 1 through 4, but it is important to anchor these to the experimental methods on which the conclusions from such investigations must ultimately be based. This is the aim of the present chapter.

ABIOTIC REACTIONS Although these reactions have already been illustrated, a few brief comments on experimental aspects should be given. For chemical reactions, few details are needed since their design follows those used in traditional preparative organic chemistry integrated with measurements of rates by methods that are widely used in physical organic chemistry. Two issues are worth noting and should be taken into account: 1. Substrate concentrations that are environmentally realistic are much lower than those often used in purely chemical investigations. 2. Environmental temperatures are highly variable, and range from low summer water temperatures in northern latitudes to high temperatures on the surface of soils during periods of high insolation. Experiments on photochemical reactions and transformations have been carried out under a number of different conditions: 1. Reactions in liquid medium can be carried out by illumination with radiation of the relevant wavelengths generally in quartz vessels of various configurations. 2. Gas-phase reactions have been carried out in 160 mL quartz vessels, and the products analyzed online by mass spectrometry (Brubaker and Hites 1998). Hydroxyl radicals were produced by photolysis of ozone in the presence of water: O3 + hν → O(1D) + O2; O(1D) + H2O → 2OH 3. Reactions that simulate tropospheric conditions have been carried out in Teflon bags with volumes of ca. 6 m3 fitted with sampling ports for introduction of reactants and substrates, and removal of samples for analysis. Substrates can be added in the gas phase or as aerosols that form a surface film. The primary reactants are the hydroxyl and nitrate radicals, and ozone. These must be prepared before use by reactions (a) to (c). a. Hydroxyl radicals by photolysis of methyl nitrite: CH3–ONO + hν → CH3O + NO CH3O + O2 → CH2O + HO2 HO2 + NO → OH + NO2 245

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NO is added to inhibit photolysis of NO2 that would produce O3 and NO3 radicals: NO + hν → NO + O(3P) O(3P) + O2 → O3 O3 + NO2 → NO3 + O2 b. Nitrate radicals are prepared by photolysis of dinitrogen pentoxide in the dark: N2O5 → NO3 + NO2 c. Ozone by a corona discharge in O2 Products from the reactions are collected on Tenax cartridges, and the analytes desorbed by heating, or on polyurethane form plugs from which the analytes can be recovered by elution with a suitable organic solvent. 4. Outdoor reactors (EUROPHORE) that use atmospheric illumination have been constructed in Valencia, Spain using hemispherical chambers with a volume of 170 m3 covered with ethene–propene foil that is suitably transparent at the appropriate wavelengths. It is possible to carry out reactions throughout the day, and by using differential optical spectroscopy with a path length of 326.8 m and Fourier transform spectroscopy with a path length of 128 m built into the system, possible artifacts from sampling or treatment for analysis can be minimized. In one experiment (Klotz et al. 1998) on the reaction of toluene, which is a major atmospheric contaminant, a range of concentrations of Nox was used that are observed in the lower atmosphere, and ozone concentrations at the beginning of the reactions were low. It was therefore possible to provide an unequivocal description of the photochemical reactions that were involved.

MICROBIAL REACTIONS The study of microbial metabolism is a relatively young discipline, not much more than 100 years old, like organic chemistry on which it critically depends. And the cardinal experimental procedures for isolation of microorganisms for studies on the metabolism of xenobiotics remain those of elective enrichment pioneered by Winogradsky, Beijerinck, Kluyver, van Niel, and their successors, backed up by the use of pure cultures using procedures developed by Koch in his classic investigations on anthrax. The major innovation has been the development of generally applicable procedures for the isolation of strictly anaerobic organisms that were introduced by Hungate (1969). There have, however, been major developments in methods for the elucidation of metabolic pathways that are discussed in later sections: 1. Availability of isotopically labeled compounds—in particular 14C, 13C, 18O2, 18OH2, and 15N. 2. Application of genetic procedures and molecular biology. 3. Application of modern analytical procedures and physical methods for structure determination. In studies of microbial metabolism, the advantages resulting from the requirement for only extremely small quantities of material needed for gas- and liquidchromatographic quantification, coupled to mass spectrometric identification, can hardly be overestimated.

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There has, at the same time, been an increasing realization of the role of populations and the intrinsic limitations in enrichment. These are noted later in this chapter, and procedures for their examination are discussed in Chapter 13. In the following sections, an attempt is made to provide a critical outline of experimental aspects of investigations directed to biodegradation and biotransformation, with particular emphasis on outstanding issues to which sufficient attention has not always been paid, and which have not therefore received ultimate resolution. Before proceeding further, it is desirable to define clearly some operational terms: Mineralization is conventionally used for the aerobic degradation of a compound to CO2 and H2O. Ready biodegradability refers to the situation in which the test compound is totally degraded (under aerobic conditions to CO2, H2O, etc.) within the time span of a standardized test usually lasting 5, 7, or 28 d. Inherent biodegradability is applied when the compound may be degraded, though not under the standard conditions generally used: their degradation may require, for example, preexposure to the xenobiotic. Recalcitrance is a valuable concept (Alexander 1975) that has been applied to compounds that have not been demonstrably degraded under the conditions used for their examination. Biotransformation is applied to situations in which even though degradation is not achieved, minor structural modifications of the test compound have occurred. Rigid boundaries between these terms should not, however, be drawn, since all of them are operational rather than absolute.

STORAGE OF SAMPLES In both field investigations and laboratory experiments, it may be necessary to store samples before examination for biodegradation of the contaminants. For laboratory experiments there is little choice except freezing. There has been conflicting advice for soil samples. Although the ISO Standard ISO/TC/190/SC 2 does not recommend freezing, for samples from three horizons of the unsaturated zone of a sandy soil, it has been shown that rates of degradation by the endogenous organisms of both 4-chloro-2-methylphenoxyacetic acid (MCPA) and metribuzin (4-amino-6-tertbutyl-3-methylthio-1,2,4-triazin-5-(4H)-one were unaffected by freezing the samples (Mortensen and Jacobsen 2004). In contrast, storage under conditions of drying followed by addition of water to the original content produced substantial changes: there were significant differences in the population of archaea and bacteria, and degradation of metalaxyl-M and lufenuron were delayed by factors up to 6 (Pesaro et al. 2004).

DETERMINATION OF READY BIODEGRADABILITY Because of the central role that estimates of biodegradability play in environmental impact assessments, a great deal of effort has been devoted to developing standardized test procedures (Gerike and Fischer 1981). In spite of this, conventional tests for biodegradability under aerobic conditions retain some questionable, or even undesirable, features from an environmental point of view. Attention is therefore drawn to two valuable critiques of widely used procedures (Howard and Banerjee 1984; Battersby 1990). A methodology based on continuous measurement of CO2 evolution by measuring conductivity provides continuous measurements in an essentially closed system (Strotmann et al. 2004). This makes it possible to evaluate substrates that are poorly soluble in water or volatile. Comparison with other widely used systems showed that it was both accurate and reliable. Some of the important issues in the design of such tests are therefore only briefly summarized here.

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The Inoculum For assessment of biodegradation in freshwater systems that have been most extensively examined, the inoculum is often taken from municipal sewage treatment plants and is therefore dominated by microorganisms that have been subjected to selection primarily for their ability to use readily degraded substrates. These organisms are clearly valuable in evaluating the persistence of compounds that might be poorly degraded in such treatment systems, and might therefore be discharged unaltered into the environment. They are, however, not necessarily equally suited to investigations on the degradability of possibly recalcitrant substances in natural ecosystems that are dominated by microorganisms adapted to a different environment. Concentration of the Substrate In conventional assay systems, substrate concentrations are generally used at levels appropriate to measurements of the uptake of oxygen or the evolution of carbon dioxide, and these concentrations greatly exceed those likely to be encountered in receiving waters; attention has been already been drawn to this issue in Chapter 4. Application to poorly water-soluble (De Morsier et al. 1987) or to volatile substrates may be difficult or even impossible due to the need for the high substrate concentrations. In addition, high substrate concentrations may be toxic to the test organism and thereby provide false-negative results. Use of 14C-labeled substrates and measurement of the evolution of 14CO2 enables very much lower substrate concentrations to be used, and this is particularly important if the contribution of marine oligotrophic bacteria is to be evaluated. This has been applied to populations of marine bacteria using 14C-labeled acetate, aniline, 4-nitrophenol, desmethyl, methylparathion, and 4-chloroaniline (Nyholm et al. 1992). The wider application of the use of isotopes is limited primarily by the availability of labeled substrates, though an increasingly wider range of industrially important compounds including agrochemicals is becoming commercially available. Endpoints For aerobic degradation, uptake of oxygen or the evolution of carbon dioxide is most widely used. Use of the concentration of dissolved organic carbon may present technical problems when particulate matter is present, though analysis of dissolved inorganic carbon in a closed system has been advocated (Birch and Fletcher 1991), and may simultaneously overcome problems with poorly soluble or volatile compounds. For anaerobic degradation, advantage has been taken of methane production (Birch et al. 1989; Battersby and Wilson 1989). Whereas this may be valuable in the context of municipal sewage treatment plants, it is more questionable whether this is generally a valid parameter in investigations concerned with anaerobic degradation in natural ecosystems in view of the extensive evidence for anaerobic degradation by nonmethanogenic bacteria such as sulfate-reducing anaerobes.

DESIGN OF EXPERIMENTS ON INHERENT BIODEGRADABILITY Probably most investigations have been carried out in conventional batch cultures, but attention should be drawn to an attractive and flexible procedure using a cyclone fermentor (Liu et al. 1981). Metabolic Limitations In the most widely applied procedures, the test system is restricted in flexibility by the salinity and pH requirements of the test organisms, but probably the most serious limitation of these test systems

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is that no account is generally taken of biotransformation reactions, nor is identification of their products routinely attempted. Examples where this has been carried out include the following: 1. In a study of the degradation of sodium dodecyltriethoxy sulfate under mixed-culture die-away conditions using acclimated cultures (Griffiths et al. 1986), the metabolites were identified and the kinetics of their synthesis compared with the degradation pathways elucidated in investigations using pure cultures (Hales et al. 1982, 1986). 2. The biodegradation of branched-chain alkanol ethoxyethylates was carried out by the standard OECD confirmatory tests and the metabolites fractionated after solid-phase extraction. The structures of the metabolites were determined by electrospray mass spectrometry and this made it possible to derive a scheme for the partial degradation of the compounds (Di Corcia et al. 1998). These procedures could advantageously receive general application. Application to Marine Systems In view of the substantial quantities of xenobiotics that enter the marine environment, surprisingly limited effort has been directed to this problem. The degradation of several structurally diverse substrates including nitrilotriacetate, 3-methylphenol and some chlorobenzenes was evaluated from the rates of incorporation of 14C-labeled substrates into biomass and production of 14CO2. These were used to evaluate the differences between freshwater, estuarine, and marine environments and revealed the difficulty of correlating rates with characteristics of the microbial community (Bartholomew and Pfaender 1983). Recent studies have used both dissolved organic carbon and oxygen uptake as parameters (Nyholm and Kristensen 1992), or analysis for specific 14C-labeled compounds at low substrate concentration, which are proposed as “simulation tests” (Nyholm et al. 1992). Both these tests used the indigenous organisms present in seawater and thereby provided a valuable degree of relevance even though they inevitably encountered the variability in the nutritional status—particularly for organic carbon in seawater. Attention is drawn to oligotrophic marine ultramicrobacteria that are of undoubted importance in oceanic systems. There are, however, important aspects of their isolation that should be appreciated, including the extent to which their physiology may be altered during maintenance under severe carbon limitation in the laboratory (Schut et al. 1993). Biodegradation of xenobiotics by these organisms has not attracted the attention it merits, and it is hazardous to extrapolate results from freshwater or brackish-water systems to marine systems since all the factors noted above coupled to the low temperatures that characterize ocean water introduce complexities that remain to be systematically investigated. Attention is drawn to the extensive metabolic potential of anaerobic sulfate-reducing bacteria isolated from marine muds and sediments. Examples are provided throughout in Chapters 7 through 9. Isolation and Elective Enrichment It is only seldom that it has been possible to obtain bacteria with a desired metabolic capability directly from natural habitats. Almost always large numbers of other organisms are present, so that some form of selection or enrichment is generally adopted before metabolic studies are attempted. Use of antibiotics or even more drastic procedures using alkali or hypochlorite have been used only infrequently except for the isolation of pathogenic bacteria such as Mycobacterium tuberculosis. Valuable results have been obtained from experiments using metabolically stable mixed cultures, which overcome to a large extent the problems of repeatability. Probably most metabolic studies on xenobiotics have, however, been carried with pure cultures of organisms. With the possible exception of anaerobic bacteria, and in particular methanogens, only a limited range of these organisms have been isolated from samples of municipal sewage sludge. Most have been obtained after elective enrichment of natural samples of water, sediments, or soils. This methodology was developed

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by Beijerinck and Winogradsky, and has been extensively exploited in the pioneering investigations carried out by the Delft school and their successors over many years. In the present context, one of its particularly attractive features is the inherent degree of environmental realism introduced by its application, and the flexibility whereby virtually any environmental condition can be reproduced must be considered as one of the most attractive features of this procedure. There is extensive evidence for the existence of enrichment in the natural environment and a number of examples illustrating its operation have been given in Chapter 4. Possibly the last word should be left to one of the pioneers of its application: “But once an elective culture method for a particular microbe is available, it may safely be concluded that this organism will also be found in nature under conditions corresponding in detail to those of the culture, and that it will carry out the same transformations.” (Van Niel 1955)

Enrichment may, however, introduce a bias, not only in the organisms but also in their degradative genes (Marchesi and Weightman 2003). In addition, there are many organisms that have not hitherto been cultivated (Stevenson et al. 2004), and these almost certainly play an important role in degradation. Natural enrichment certainly takes place in environments that have been chronically exposed to contaminants and is an important mechanism for the loss of agrochemicals. Degradation of contaminants may, however, be observed in bacteria that have been isolated from pristine environments with no established exposure to the contaminants (e.g., Kitagawa et al. 2002; Sánchez et al. 2004). Procedures that have been used to investigate the role of natural populations without isolation of the relevant organisms are discussed later in Chapter 13. General Procedures In its simplest form, the procedure consists of the following steps: a. Elective enrichment of the microorganisms in an environmental sample by growth at the expense of a single compound serving as the sole source of carbon and energy. b. Successive transfer into fresh medium after growth has occurred. c. Isolation of the appropriate organisms, generally by plating on solid media or by dilution. Some of the experimental details are briefly described in the following paragraphs. The three successive stages used in isolating the desired organisms are outlined first, followed by a more extensive discussion of media. 1. An appropriate mineral medium supplemented with the organic compound that is to be studied is inoculated with a sample of water, soil, or sediment. In studies of the environmental fate of a xenobiotic in a specific ecosystem, samples are generally taken from the area putatively contaminated with the given compound so that a degree of environmental relevance is automatically incorporated. Attention has, in addition been directed to pristine environments, and the issues of adaptation or preexposure have already been discussed. If the test compound is to serve as a source of sulfur, nitrogen, or phosphorus, these elements will necessarily be omitted from the medium, and an appropriate carbon source supplied either by the xenobiotic under investigation, or by another substrate that is added. In such experiments, glassware must be scrupulously cleaned to remove interfering traces of, for example, detergents since these may contain residues of all of these nutrients and could therefore compromise the outcome of the experiment. The metabolism of phosphonates and sulfonates, and of sulfate and phosphate esters, may in addition, be inducible only in the absence of inorganic sources of phosphorus or sulfur. Although aromatic sulfonates have been used as sources of sulfur, the pathway for their aerobic degradation differs

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from that when they are used as sources of carbon and energy. The following are given as illustration of enrichments designed to obtain organisms using organic compounds as sources of nitrogen, sulfur, or phosphorus—although they are not necessarily able to use the test substrates as carbon sources: a. Nitrophenols as nitrogen source using succinate as carbon source (Bruhn et al. 1987). b. 2-Chloro-4,6-diamino-1,3,5-triazine as nitrogen source and lactate as carbon source (Grossenbacher et al. 1984), or 2-chloro-4-aminoethyl-6-amino-1,3,5-triazine as nitrogen source and glycerol as carbon source (Cook and Hütter 1984). c. Arylsulfonic acids as sulfur source and succinate or glycerol as carbon sources (Zürrer et al. 1987). d. 2-(1-Methylethyl)amino-4-hydroxy-6-methylthio-1,3,5-triazine as sulfur source and glucose as carbon source (Cook and Hütter 1982). e. Glyphosate (N-phosphonomethylglycine) as phosphorus source and glucose as carbon source (Talbot et al. 1984). These procedures may clearly result in the dominance of organisms that carry out only biotransformation of the xenobiotic, although the biodegradation of many of these compounds has also been demonstrated using the same or other organisms. 2. The cultures are then incubated under relevant conditions of temperature, pH, and oxygen concentration, and after growth has occurred, successive transfer to fresh medium is carried out; interfering particulate matter will thereby be removed by dilution, and a culture suitable for isolation may be obtained. Incubation is generally carried out in the dark, though for the isolation of phototrophic organisms, illumination at suitable wavelength and intensity must obviously be supplied. There are no rigid rules on how many transfers should be carried out, but especially for anaerobic organisms, sufficient time should elapse between transfers to allow growth of these often slow-growing organisms. Enrichment may therefore take up to a year or even longer. 3. After metabolically stable cultures have been obtained, pure cultures of the relevant organisms may then be obtained by any of three basic procedures: a. By preparing serial dilution of the culture in a suitable buffer medium and spreading portions onto solid media (agar plates or roll tubes for anaerobic bacteria) containing the organic compound as source of carbon, sulfur, nitrogen, or phosphorus. The plates (or tubes) are then incubated under appropriate conditions. After growth has taken place, single colonies are then selected and pure cultures obtained by repeated restreaking on the original defined medium. Use of complex media or substrate analogs may introduce serious ambiguities since overgrowth of unwanted, rapidly growing organisms may occur. Considerable difficulty may be experienced when “spreading” organisms conceal the desired organism and this may make the isolation of single colonies a tedious procedure. b. Serial dilution may be carried out in a defined liquid growth medium and the dilutions incubated under suitable conditions; successive transfers are then made from the highest dilution showing growth. This experimentally tedious procedure may indeed be obligatory for organisms such as Thermomicrobium fosteri (a name no longer accepted by Zarilla and Perry 1984), which are unable to produce colonies on agar plates (Phillips and Perry 1976). It has been quite extensively used for anaerobic bacteria in which the liquid medium is replaced by a soft agar medium. A cyclic procedure involving an additional plating step was used to reduce the complexity of the original population and to isolate an organism utilizing high concentrations of 2,4,6-trichlorophenol (Maltseva and Oriel 1997). c. Mechanical methods may be used for the isolation of filamentous organisms. For example, washing on the surface of membrane filters, or micromanipulation on the surface

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of agar medium (Skerman 1968) has occasionally been employed, and an ingenious procedure, which uses an electron microscope grid for preliminary removal of other organisms, has been described for use under anaerobic conditions (Widdel 1983). Lack of repeatability of the results of metabolic studies using laboratory strains that have been maintained by repeated transfer for long periods under nonselective conditions may be encountered. These strains may no longer retain their original metabolic capabilities, and this may be particularly prevalent when the strains carry catabolic plasmids that may have been lost under nonselective conditions. For these reasons, strains should be maintained in the presence of a cryoprotectant such as glycerol at low temperatures (−70°C or in liquid N2) as soon as possible after isolation. Freeze-drying is also widely adopted, and is recommended. In some cases, difficulty may be experienced in isolating pure cultures with the desired metabolic capability, and the mixed enrichment cultures or consortia must be used for further studies. For example, although an enrichment culture effectively degraded atrazine (2-chloro-4-ethylamine6-isopropylamine-1,3,5-triazine), none of the 200 pure cultures isolated from this were able to use the substrate as nitrogen source (Mandelbaum et al. 1993). Two (or more) organisms may cooperate in the degradation of the substrate. Considerable effort may then be required to determine the appropriate combination of organisms. For example, an enrichment using 4-chloroacetophenone yielded eight pure strains although none of these could degrade the substrate in pure culture. All the pair combinations were then analyzed, and this revealed that the degradation of the substrate was accomplished by strains of an Arthrobacter sp. and a Micrococcus sp. These were then used to elucidate the metabolic pathway (Havel and Reineke 1993). Attention is drawn to the fact that, in this case, isolation of the pure strains from the enrichment culture was carried out using a complex medium under nonselective conditions. These batch procedures for enrichment and successive transfer may be replaced by the use of continuous culture. This may be particularly attractive when the test compound is toxic, when it is poorly soluble in water, or where the investigations are directed to substrate concentrations so low that clearly visible growth is not to be expected. These problems remain, however, for subsequent isolation of the relevant organisms. One considerable problem in long-term use arises from growth in the tubing of the pump system that is used to administer the medium and should be renewed periodically. Basal Media The choice of appropriate basal media is of cardinal importance and a number of important practical considerations should be taken into account. Mineral Media A plethora of basal media for the growth of freshwater organisms has been formulated. These may, however, differ significantly, particularly in the concentrations of phosphate, while for anaerobic bacteria the inclusion of bicarbonate and a suitable reductant is a standard practice. Numerous examples of suitable media have been collected in The Prokaryotes (Balows et al. 1992). Clearly, if the organic substrate is to serve as a source of nitrogen, sulfur or phosphorus, these elements must be omitted from the basal medium. Otherwise these inorganic nutrient requirements will generally be supplied by the following: sulfur (generally as sulfate except for organisms such as chlorobia, which require reduced sulfur as S2−); nitrogen (generally as ammonium or nitrate except for N2-fixing organisms); phosphorus (as phosphate); Mg2+, Ca2+, and lower concentrations of Na+ and K+. For marine organisms, the basal medium is constructed to resemble natural seawater in the concentrations of Na+, K+, Mg2+, Ca2+, Cl−, and SO42−, while glycerophosphate may be used as phosphorus to avoid problems with precipitation. A number of different formulations have been used (Taylor et al. 1981, Neilson 1980), though the precise composition does not seem to be critical.

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The inclusion of nitrate may lead to various complications, which have been discussed in Chapter 2. In addition, the nitrogen status of the growth medium determines the levels of lignin peroxidases and manganese-dependent peroxidases that are synthesized in Phanerochaete chrysosporium. The role of Mn concentration is noted later and in Chapter 3, Part 5. It is important to underscore the fact that carbon dioxide is required for the growth of not just strictly phototrophic and lithotrophic organisms. Many organisms that are heterotrophic have an obligate requirement for carbon dioxide for their growth. Illustrative examples include the following: 1. Anaerobic bacteria such as the acetogens, methanogens and the propionic bacteria—the medium is generally supplemented with bicarbonate. 2. Aerobic bacteria include the following: a. Degradation of propane (MacMichael and Brown 1987) and the branched-chain hydrocarbon 2,6-dimethyloct-2-ene (Fall et al. 1979) b. Metabolism of epoxides by Xanthobacter sp. strain Py2 (Ensign et al. 1998) c. Oxidation of carbon monoxide (Meyer and Schlegel 1983) d. Degradation of acetone by both aerobic and anaerobic bacteria. Trace Elements Many enzymes and coenzymes contain metals (Chapter 3, Part 5), and these elements—Zn, Cu, Fe, Mn, Co, Ni, and Mo—are generally provided at low concentrations. These are provided as the cation salts and for Mo as molybdate. A large number of different formulations of trace elements have been published. The A4 formulation (Arnon 1938) supplemented with Co2+ and MoO42− has been widely used, or one of the SL series of formulations developed by Pfennig and his coworkers particularly for the cultivation of anaerobic bacteria. SL 9 is typical of the later series of formulations (Tschech and Pfennig 1984). Somewhat conflicting views exist on the possibly deleterious effects resulting from the incorporation of complexing agents, particularly ETDA, so that their concentration should probably be kept to a minimum. Anaerobic bacteria are more fastidious in their trace-metal requirements, and selenium as selenite and tungsten as tungstate are routinely added (Tschech and Pfennig 1984). Selenium is required for the synthesis of active xanthine dehydrogenase in purine-fermenting clostridia (Wagner and Andreesen 1979), and of formate dehydrogenase in a number of organisms including methanogens (Jones and Stadtman 1981) and clostridia (Yamamoto et al. 1983). Tungsten is required for the synthesis of a range of enzymes necessary for the growth of various methanogens (Zellner et al. 1987; Winter et al. 1984), an anaerobic cellulose-degrading bacterium (Taya et al. 1985), and for the synthesis of various enzymes in anaerobic hyperthermophilic bacteria. Further important examples that are dealt with in later sections include 1. The carboxylic acid reductase in acetogenic clostridia such as Clostridium thermoaceticum (White et al. 1989; Strobl et al. 1992). 2. The benzylviologen-linked aldehyde oxidoreductase in Desulfovibrio gigas grown with ethanol (Hensgens et al. 1995) and the corresponding enzyme in Desulfovibrio simplex (Zellner and Jargon 1997). For the latter, it was suggested that the flavins FMN or FAD were the natural cofactors. 3. The acetylene hydratase of Pelobacter acetylenicus (Rosner and Schink 1995). 4. Tungsten may be incorporated into some proteins of purinolytic clostridia (Wagner and Andreesen 1987), and into formylmethanofuran dehydrogenase in Methanobacterium thermoautotrophicum (Bertram et al. 1994). 5. Ferredoxin reductases in Pyrococcus furiosus including aldehyde ferredoxin reductase, glyceraldehyde-3-phosphate ferredoxin oxidoreductase, and formaldehyde ferredoxin reductase (Roy et al. 1999).

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It should be noted, however, that whereas the production and activity of some enzymes in Thermotoga maritima was stimulated by tungstate, the purified enzymes apparently did not contain tungsten. For the sake of completeness, it is worth noting the role of vanadium in the synthesis of alternative nitrogenase systems by several aerobic bacteria, particularly those within the genus Azotobacter (Fallik et al. 1991); low yields of hydrazine are produced, and this suggests a different affinity of the nitrogenase for the N2 substrate (Dilworth and Eady 1991). Control of pH Although media are often buffered by the inclusion of phosphate, excessive concentrations should be avoided since precipitation of insoluble Ca and Mg phosphates may occur during sterilization by autoclaving. A number of organic buffers have been used effectively in several applications. In studies employing organic phosphorus compounds as sources of phosphorus, phosphate buffer has been replaced by, for example, HEPES (Cook et al. 1978), and MOPS has been incorporated into media for growth of Enterobacteriaceae (Neidhardt et al. 1974). For media requiring low pH, MES has been used in a medium with an extremely low phosphate concentration (Angle et al. 1991), and TRIS has been incorporated into media for growth of marine bacteria (Taylor et al. 1981). For all of these, it should be established that metabolic complications do not arise as a result of the ability of the organisms to use the buffer as sources of nitrogen or sulfur. It is also possible for the buffer to react with metabolic intermediates, and this is illustrated by the isolation of a compound produced by reaction of the carbon atom (formally CO) of CCl4 with HEPES (Lewis and Crawford 1995). Metal chelating agents such as NTA or ETDA may be used, although their concentrations should be kept to a minimum in view of their potential toxicity: NTA must obviously be omitted in studies of utilization of organic nitrogen as N-source since both NTA and its metal complexes are apparently quite readily degraded (Firestone and Tiedje 1975). Vitamins Vitamins such as thiamin, biotin, and vitamin B12 are often added. Once again, the requirements of anaerobes are somewhat greater, and a more extensive range of vitamins that includes pantothenate, folate, and nicotinate is generally employed. In some cases, additions of low concentrations of peptones, yeast extract, casamino acids or rumen fluid may be used, though in higher concentrations, metabolic ambiguities may be introduced since these compounds may serve as additional carbon sources. Sterilization Mineral–basal media may be sterilized by autoclaving, but for almost all organic compounds that are used as sources of C, N, S, or P, it is probably better to prepare concentrated stock solutions and sterilize these by filtration, generally using 0.2 µm cellulose nitrate or cellulose acetate filters. The same applies to solutions of vitamins, and to solutions of bicarbonate and sulfide that are components of many media used for anaerobic bacteria. Metal Concentration in Metabolism These requirements should be assessed in the context of metalloenzymes, which are discussed in this section. Iron Although Fe is required as a trace element, its uptake is critically regulated since excess leads to the generation of toxic hydroxyl radicals, and complex interactions involving Fe(II) and Fe(III) exist within the cell (Touati et al. 1995). The role of Fe(III)-complexing siderophores in maintaining homeostasis has been noted in Chapter 3, Part 5. Details of the role of Fe and its relation to the

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generation of toxic hydroxyl radicals have been explored by analysis of a strain of Escherichia coli and a mutant strain lacking both Fe and Mn superoxide dismutase. The mutant strain showed a marked increase in the hydroxyl radical after exposure to H2O2 (McCormick et al. 1998): preincubation with an Fe chelator inhibited this difference, and redox-active Fe defined as EPR-detectable ascorbyl radicals was greater in the mutant than in the wild-type strain. Iron may also play a subtler role in determining the biodegradability of a substrate that forms complexes with Fe. Two examples are used as illustration of probably different underlying reasons for this result: a. A strain of Agrobacterium sp. was able to degrade ferric EDTA, though not the free compound (Lauff et al. 1990). This may be due either to the adverse effect of free EDTA on the cells or the inability of the cells to transport the free compound. The former is supported by the established sensitivity (Wilkinson 1968) of some Gram-negative organisms to ETDA, and the increased surface permeability in enteric organisms exposed to EDTA (Leive 1968). These results are consistent with the requirement for high concentrations of Ca2+ in the enrichment medium used for isolating a mixed culture capable of degrading EDTA (Nörtemann 1992). In contrast, although the degradative enzymes have been purified and characterized from cells of a pure culture grown with Mg ETDA, the enzyme complex was unable to use Fe EDTA as a substrate. (Witschel et al. 1997). This is consistent with the results using an unenriched culture and 14C-labeled EDTA (Allard et al. 1996). In the aquatic environment with low concentrations of other complexing cations, degradation of EDTA—and probably DTPA—is accomplished primarily by photolysis of the Fe complex (Kari and Giger 1995). b. A strain of Pseudomonas fluorescens biovar II is able to degrade citrate whose metabolism requires access to the hydroxyl group. This group is, however, implicated in the tridentate ligand with Fe(II) so that this complex is resistant to degradation in contrast to the bidentate ligand with Fe(III) that has a free hydroxyl group and is readily degraded (Francis and Dodge 1993). Similarly, the bidentate complexes containing Fe(III), Ni, and Zn are readily degraded by Pseudomonas fluorescens, in contrast to the tridentate complexes containing Cd, Cu, and U that are not degraded (Joshi-Tope and Francis 1995). These results may be viewed in the wider context of interactions between potential ligands of multifunctional xenobiotics and metal cations in aquatic environments and the subtle effects of the oxidation level of cations such as Fe. The Fe status of a bacterial culture has an important influence on synthesis of the redox systems of the cell since many of the electron transport proteins contain Fe. This is not generally evaluated systematically, although the degradation of tetrachloromethane by a strain of Pseudomonas sp. under denitrifying conditions clearly illustrated the adverse effect of Fe on the biotransformation of the substrate (Lewis and Crawford 1993; Tatara et al. 1993). This possibility should therefore be taken into account in the application of such organisms to bioremediation programs. Manganese The role of manganese concentration has seldom been explicitly examined in the context of biodegradation. A few extradiol dioxygenases contain Mn in place of Fe, and it is essential for the growth of the purple nonsulfur anaerobic phototrophs Rhodospirillum rubrum and Rhodopseudomonas capsulata during growth with N2, though not with glutamate (Yoch 1979). It plays an essential role in the metabolic capability of the white-rot fungus Phanaerochaete chrysosporium. This organism produces two groups of peroxidases during secondary metabolism—lignin peroxidases and manganese-dependent peroxidases. Both are synthesized when only low levels of Mn (II) are present in the growth medium, whereas high concentrations of Mn result in repression of the synthesis of the lignin peroxidases and an enhanced synthesis of manganese-dependent peroxidases (Bonnarme and Jeffries 1990; Brown et al. 1990). Experiments with a nitrogen-deregulated mutant have shown

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that N regulation of both these groups of peroxidases is independent of Mn(II) regulation (Van der Woude et al. 1993). The unusual manganese-dependent dioxygenase (Hatta et al. 2003) would also necessitate adequate concentrations. Copper In Methylosinus trichosporium OB3b, which expresses a particulate monooxygenase, the concentration of copper plays a significant role in the kinetic parameters for the consumption of both methane and trichloroethene (Sontoh and Semrau 1998). For methane, Vmax decreased from 300 to 82 when the concentration of Cu increased from 2.5 to 20 µM, and Ks decreased from 62 to 8.3 under these conditions. For trichloroethene, Vmax and Ks were unmeasurable at Cu concentrations of 2.5 µM even in the presence of formate, but were 4.1 and 7.9 for concentrations of 20 µM in the presence of formate. Although copper is an unusual component of dioxygenases, it is a component of the dioxygenase from Aspergillus japonicus, which brings about fission of the C ring of quercitin with formation of carbon monoxide (Steiner et al. 2002). It is also a component of oxalate decarboxylase in Bacillus subtilis, which converts oxalate into formate and CO2 (Tanner et al. 2001). Other Metals Heavy-metal cations and oxyanions are generally toxic to bacteria although resistance may be induced by various mechanisms after exposure. Attention is drawn to an unusual example in which Al3+ may be significant, since the catechol 1,2-dioxygenase and 3,4-dihydroxybenzoate (protocatechuate) 3,4-dioxygenase that are involved in the metabolism of benzoate by strains of Rhizobium trifolii are highly sensitive to inhibition by Al3+ (Chen et al. 1985). Redox Potential of Media Cultivation of strictly anaerobic organisms requires not only that the medium be oxygen-free, but also that the redox potential of the medium be compatible with that required by the organisms. This may be accomplished by addition of reducing agents such as sulfide, dithionite, titanium(III) citrate, or titanium(III) nitrilotriacetate. Any of these may, however, be toxic so that only low concentrations should be employed. Attention has been drawn to the fact that titanium(III) citrate–reduced medium may be inhibitory to bacteria during initial isolation (Wachenheim and Hespell 1984).

ORGANIC SUBSTRATES Although organic substrates such as carboxylic acids are thermally stable and may be sterilized with the basal media, many others including, for example, carbohydrates, esters, or amides are better prepared as concentrated stock solutions, sterilized by filtration through 0.22 µm filters and added to the sterile basal medium. Toxic Compounds A major problem arises when the desired organic substrates are poorly soluble in water, highly volatile, or toxic. For example, it has been found convenient to supply toxic substrates such as hydrogen cyanide in the gas phase (Harris and Knowles 1983). Attention should be directed to the requirement of many organisms for CO2, which has already been noted, and is discussed later in this section. Serious problems may arise for substrates that are too toxic to be added in the free state at concentrations sufficient for growth. For example, the toxicity of long-chain aliphatic compounds with low water solubility has been examined in yeasts (Gill and Ratledge 1972). The principles that emerged have been applied to circumvent toxicity in liquid media by several strategies: a. Application of toluene in an inert hydrophobic carrier (Rabus et al. 1993). b. The water-soluble hexadecyltrimethylammonium chloride adsorbed on silica (Van Ginkel et al. 1992).

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c. A gas-phase delivery system using tetrachloroethene sorbed on beads of Tenax for continuous supply to an anaerobic Desulfuromonas sp. strain BB1 (Brennan and Sanford 2002). d. Dibutyl phthalate as a cosolvent for the toxic cyclohexane epoxide has been examined, though in this case, the essential problem was the susceptibility of the substrate to hydrolysis to the diol that was more readily degradable (Carter and Leak 1995). e. Bio-Sep beads amended with 13C-labeled monocyclic arenes to assess in situ biodegradation potential (Geyer et al. 2005). Methods that have been applied to the preparation of solid media are described later. Volatile Compounds Gaseous or highly volatile substrates present a problem that may be overcome by the use of enclosed systems such as desiccators or sealed ampoules, or by bottles capped with teflon-lined crimp caps and inverted during incubation. Attention is drawn, however, to the permeability of many types of rubber sealing to organic compounds. This procedure has been employed for 4-chloroacetophenone (Havel and Reineke 1993), since the 4-chlorophenol produced was toxic to one of the components of the consortium. In this example, low concentrations of the toxic intermediate could also be maintained by adding gelatin to the medium, and this procedure facilitated the growth of one of the components. It should be noted that, particularly in sealed systems, it is important to satisfy the obligate requirement of many organisms for CO2. The result of enrichments may sometimes depend upon whether the substrate is applied in the vapor phase or in the aqueous phase; for example, enrichments with α-pinene in the vapor phase yielded predominantly Gram-positive organisms in contrast to the Gram-negative organisms obtained when the substrate was added to the liquid medium. This may plausibly be related to the greater sensitivity of Gram-positive organisms that are exposed to only low substrate concentrations in the vapor phase (Griffiths et al. 1987).

SOLID MEDIA Although solid media have been prepared from silica gels these have not been widely used. Agar for preparing solid medium should be of the highest quality, and as free as possible from alternative carbon sources. It is generally preferable to autoclave agar separately from the mineral base; both are prepared at double the final concentration and mixed after autoclaving (Stanier et al. 1966). The problem of preparing plates for testing the metabolic capacity of substrates that are only poorly soluble is a serious one for which no universal solutions are available. Some of the techniques that have been advocated include the following: 1. Liquid hydrocarbons have been adsorbed on silica powder and dispersed in the agar medium. The silica may be autoclaved (Baruah et al. 1967), though this may be avoided by sterilizing the silica by heating, and carrying out the sorption and removal of solvents such as acetone or dichloromethane under sterile conditions. 2. Solutions of the substrate have been prepared, for example, in acetone or diethyl ether, and added to or spread over the surface of agar plates either before or after inoculation (Sylvestre 1980; Shiaris and Cooney 1983). 3. For compounds that are sufficiently volatile such as benzene, toluene, or naphthalene, the substrate may be contained in a tube placed above the level of the medium (Claus and Walker 1964), or on the lid of a Petri dish (Söhngen 1913). For benzene and toluene this also obviates problems with toxicity since the organisms are exposed to only low concentrations of the substrate. 4. A solution of the substrate in ethanol may be mixed with the bacterial suspension in agarose and poured over agar plates of the base medium (Bogardt and Hemmingsen 1992).

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This is the general procedure used with top agar in the Ames test (Maron and Ames 1983), although dimethyl sulfoxide is generally used as the water-miscible solvent. 5. Chthoniobacter flavus has been obtained in pure culture (Sangwan et al. 2004), which was made possible by the use of a dilute nutrient broth medium solidified with gellan gum in place of agar (Janssen et al. 2002), and this has since been extended (Sangwan et al. 2005).

GROWTH AT THE EXPENSE OF ALTERNATIVE SUBSTRATES It may be found that after enrichment, growth does not occur on plates prepared with the compound that showed satisfactory growth in liquid medium. An alternative that is worth examining is that of attempting to grow the organisms with a potential metabolite, though it should be kept in mind that organisms may be unable to utilize compounds that are clearly established metabolites of the desired substrate. Examples include the inability of fluorescent pseudomonads, which degrade aromatic compounds via cis,cis-muconate to use this as a substrate (Robert-Gero et al. 1969) or of alkane-degrading bacteria to grow with the corresponding carboxylic acids which are the first metabolites in alkane degradation (Zarilla and Perry 1984). Plausible reasons could be the lack of an effective transport system for the metabolite or the failure of the metabolite to induce the enzymes necessary for its production. For example, whereas salicylate normally induces the enzymes required for the degradation of naphthalene, this is apparently not the case for a naphthalene-degrading strain of Rhodococcus sp. (Grund et al. 1992). Use of complex media for isolating organisms after elective enrichment is, in contrast, a potentially hazardous procedure. Media that are routinely used for nonmetabolic studies in clinical laboratories generally contain high concentrations of peptones, casamino acids, yeast extract, or carbohydrates. These may provide alternative carbon sources, and their use may therefore result in only mild selection pressure for the emergence of the desired organisms: overgrowth by undesired microorganisms may then take place all too readily.

TECHNIQUES FOR ANAEROBIC BACTERIA Increased attention has been directed to the growth and isolation of anaerobic bacteria. In addition to the nutritional requirements noted above, their general requirement for CO2 should be taken into consideration. Broadly, three types of experimental procedures have been used: 1. Anaerobe jars containing a catalyst for the reaction between oxygen and hydrogen that is either added to or generated within the system. These systems have significant limitations in the kinds of experiment that can be carried out, since at some stage exposure to air cannot be avoided. In addition, many workers have experienced the unreliability of these systems, and they are not suitable for work with highly oxygen-sensitive organisms such as methanogens. 2. The classical technique (Hungate 1969) has been successfully used over many years, and incorporates a number of features designed to minimize exposure to oxygen. This procedure enables incubation to be carried out under a variety of gas atmospheres that include CO2 and H2, and is designed to produce a redox potential in media that is suitable for growth. Roll-tubes have been used instead of Petri dishes, and thereby a strictly anoxic environment may be maintained during manipulation. A modification using serum bottles has been introduced (Miller and Wolin 1974), and roll-tubes have also been successfully used for isolation of anaerobic phycomycetous fungi from rumen fluid (Joblin 1981). 3. Anaerobe chambers of varying design have achieved increasing popularity since they enable standard manipulations to be carried out under anoxic conditions. These systems maintain a gas atmosphere of N2, H2, and CO2 (generally 90:5:5), and include a heated catalyst for the maintenance of anaerobic conditions. They may employ either a glove-box design, or free access through wristbands, and enable quite sophisticated experiments to be carried out and cultures maintained over lengthy periods. These can therefore be unequivocally recommended although their maintenance costs should not be underestimated.

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DESIGN OF EXPERIMENTS ON BIODEGRADATION AND BIOTRANSFORMATION There are several essentially different kinds of experiments that may be carried out. Two are laboratory-based and two are field investigations: 1. 2. 3. 4.

Laboratory experiments using pure cultures or stable consortia Laboratory experiments using communities in microcosms simulating natural systems Field experiments in model ecosystems—pools or steams Large-scale field experiments under natural conditions

It should be clearly appreciated that the objectives of these various investigations are rather different. The first two aim at elucidating the basic facts of metabolism, the products formed, and the kinetics of their synthesis. Studies using pure cultures may ultimately be directed to studying more sophisticated aspects of the regulation and genetics of biodegradation. In contrast, the last two procedures are designed to obtain data of greater environmental relevance and may profitably—and even necessarily—draw upon the results obtained using the first two procedures. While the degree of environmental realism increases from Procedures 1 through 4, so also do the experimental difficulties and the interpretative ambiguities. All the procedures have clear advantages for specific objectives and are complementary. Indeed, it is desirable to use several of them to provide a broader perspective on the biodegradability of the substrate. There are significant differences in the control experiments that are possible in each of these systems. Before the quantifier bio- can be applied, the possibility of abiotic alteration of the substrate during incubation must be eliminated or taken into consideration. Only the first design lends itself readily to this control. For experiments using cell suspensions, the obvious controls are incubation of the substrate in the absence of cells or using autoclaved cultures. Care should be exercised in the interpretation of the results, however, since some reactions may apparently be catalyzed by cell components in purely chemical reactions. The question may then legitimately be raised whether or not these are biochemically mediated. Two examples are given as illustration of apparently chemically mediated reactions, which have been referred to in Chapter 1: 1. Dechlorination reactions of organochlorine compounds that involve corrinoids or porphyrins. 2. Reduction of aromatic nitro compounds by sulfide catalyzed by extracellular compounds excreted into the medium during growth of Streptomyces griseoflavus (Glaus et al. 1992). In experiments where relatively small volumes of sediment suspensions are employed, autoclaving may significantly alter the structure of the sediment as well as introducing possibly severe analytical difficulties. In such circumstances, there are few alternatives to incubation in the presence of toxic agents such as NaN3, which has been used at a concentration of 2g/L. There remains, of course, the possibility that azide-resistant strains could emerge during prolonged incubation, and the possible occurrence of reactions between the substrate and azide must also be taken into consideration. Only controls using inhibitors of microbial growth are possible for microcosm experiments, and these may be impractical for outdoor systems, which therefore combine and may fail to discriminate abiotic and biotic reactions.

PURE CULTURES AND STABLE CONSORTIA Different kinds of experimental procedures have been used and these should be evaluated against the background in which several organisms or several substrates are simultaneously present. There are no essential differences in the design of experiments using pure cultures and those employing metabolically stable consortia. It should be emphasized, however, that even in the latter, the experiments should be carried out under aseptic conditions; otherwise, interpretation of the results may be compromised by adventitious organisms.

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CELL GROWTH AT THE EXPENSE OF THE XENOBIOTIC In the simplest case, growth of the organism that has been isolated may be studied using the test substance that fulfills the nutritional requirement of the organism as the sole source of carbon, sulfur, nitrogen, or phosphorus. For a compound that is used as the sole source of carbon and energy, the endpoints could be growth and conversion of the substance into CO2 under aerobic conditions or growth under anaerobic conditions accompanied by, for example, production of methane or sulfide from sulfate. In some studies, however, only diminution in the concentration of the initial substrate has been demonstrated, and this alone clearly does not constitute evidence for biodegradation. Biotransformation is equally possible, and should be taken into consideration. Ideally, use may be made of radiolabeled substrates followed by identification of the labeled products. 14C, 35S, 36Cl, and 31P have been used, though the relevant labeled products may not always be available commercially, and the required synthetic expertise may not be available in all laboratories. Further comments on the use of these isotopes together with the application of the nonradioactive isotopes 13C and 19F are given in Chapter 6, Part 1. It should be noted that CO2 is incorporated not only into phototrophs and chemolithotrophs, but also into heterotrophic organisms. A review (Ensign et al. 1998) provides a brief summary of the role of CO2 in the metabolism of epoxides by Xanthobacter sp. strain Py2, and of acetone by both aerobic and anaerobic bacteria. In later chapters, considerable weight is given to the environmental significance of biotransformation and the synthesis of toxic metabolites. It is particularly desirable, therefore, to direct effort to the identification of such metabolites. This may present a substantially greater challenge than that of quantifying the original substrate for several reasons: 1. The structure of the metabolite will often be unknown, and must be predicted from knowledge of putative degradation pathways and confirmed by chemical analysis, generally including mass spectrometry and increasingly by NMR. 2. The metabolite will frequently be more polar than the initial substrate, so that specific procedures for extraction and analysis must generally be developed, and the pure compound must be available for quantification. 3. The metabolite may be transient with unknown kinetics of its formation and further degradation. In practice, there is only one really satisfactory solution: the kinetics of the transformation must be followed. The justification for this substantial increase in effort is the dividend resulting in the form of a description of the metabolic pathway including the synthesis of possibly inhibitory metabolites. An important dividend is that it may be possible to make generalizations on the degradation of other xenobiotics—structurally related or otherwise.

STABLE ENRICHMENT CULTURES Some investigations have used metabolically stable enrichment cultures that are particle free to study biotransformation. These are preferable to the use of sediment slurries even though these may be environmentally realistic, and there are several important reasons: • They avoid the ambiguities resulting from the presence of xenobiotics in the original soil or sediment. • They eliminate association of the added xenobiotic with particulate matter. • They increase the accuracy of analytical procedures. • They overcome the unresolved function of organic components in the sediment. • They make it possible to carry out reproducible experiments under clearly defined conditions. Particularly for anaerobes, however, it is often difficult to obtain pure cultures, and sediment or sludge slurries have frequently been used. Examples of studies that have used stable enrichment

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cultures that are free of sediment by successive transfer in defined media during extensive periods of time, include the following: 1. The extensive use in a series of experiments on the anaerobic dechlorination of chlorocatechols (references in Allard et al. 1994). 2. Studies on the anaerobic dechlorination of 2,3,5,6-tetrachlorobiphenyl (Cutter et al. 1998).

USE OF DENSE CELL SUSPENSIONS Dense cell suspensions have traditionally been used for experiments on the respiration of microbial cells at the expense of organic substrates, and they are equally applicable to experiments on biodegradation and biotransformation. Cells are grown in a suitable medium generally containing the test compound (or an alternative growth substrate) collected by centrifugation, washed in a buffer solution to remove remaining concentrations of the growth substrate and its metabolites, and resuspended in fresh medium before further exposure to the xenobiotic. For aerobic organisms, there are generally few experimental difficulties with three important exceptions: 1. For organisms that grow poorly in liquid medium, it may be difficult to obtain sufficient quantities of cells; cultures may then be grown on the surface of agar plates and the cells removed by scraping. 2. Organisms that have fastidious nutritional requirements may require undefined growth additives such as peptones, yeast extract, rumen fluid, or serum. Subsequent exposure to the xenobiotic may then be used to induce synthesis of the relevant catabolic enzymes. For example, the chlorophenol-degrading bacterium Mycobacterium chlorophenolicum (Rhodococcus chlorophenolicus) has been grown in media containing yeast extract or rhamnose, while exposure to pentachlorophenol was used to induce the enzymes required for the degradation of a wide range of chlorophenols (Apajalahti and Salkinoja-Salonen 1986). 3. Organisms such as actinomycetes may not produce well-suspended growth and shaking, for example, in baffled flasks or in flasks with coiled-wire inserts may be advantageous in partially overcoming this problem. For anaerobic bacteria, the same principles apply, except that additional attention must be directed to preparing the cell suspensions. Use of an anaerobic chamber in which cultures can be transferred to tightly capped centrifuge tubes is virtually obligatory, and addition of an anaerobic indicator should be used to ensure that subsequent entrance of oxygen does not take place inadvertently. On account of the inserts, screw-capped Oak Ridge tubes are convenient for centrifugation.

USE OF IMMOBILIZED CELLS Cells can be immobilized on a number of suitable matrices, confined in a reactor and the medium containing the test substrate circulated continuously. Although this methodology has been motivated by interest in biotechnology and in bioremediation technology, it is clearly applicable to laboratory experiments on biodegradation and biotransformation that could readily be carried out under sterile conditions. This procedure has been used, for example, to study the biodegradation of 4-nitrophenol (Heitkamp et al. 1990), pentachlorophenol (O´Reilly and Crawford 1989), and 6-methylquinoline (Rothenburger and Atlas 1993). A few additional comments are inserted: 1. It is possible to carry out experiments with immobilized cells in essentially nonaqueous media (Rothenburger and Atlas 1993), and this could prove an attractive strategy for compounds with limited water solubility, provided that solvents can be found that are compatible with the solubility of the substrate and the sensitivity of the cells to organic solvents. 2. Encapsulated cells have been successfully used for the commercial biosynthesis of a number of valuable compounds such as amino acids, and this technique could readily be

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adapted to investigations on biodegradation. This methodology offers the advantage that the metabolic activity of the cells can be maintained over long periods of time so that a high degree of reproducibility is guaranteed, and the stability of such systems may be particularly attractive in studies of recalcitrant compounds. 3. With appropriate experimental modifications, the various procedures could be adapted to study biodegradation under anaerobic conditions. For example, sparging with air could be replaced by the use of a gas mixture containing appropriate concentrations of CO2 and H2 in an inert gas such as N2.

APPLICATION OF CONTINUOUS CULTURE PROCEDURES These have been particularly valuable in studies using low concentrations of xenobiotics and for the isolation of consortia that have been used in elucidating metabolic interactions between the various microbial components. In many cases, consortia containing several organisms are obtained even though only a few of their members are actively involved in the metabolism of the xenobiotic. It is possible that the low substrate concentrations that have been used in these experiments favor selection for organisms that are able to take advantage of the lysis products from cells that do not play a direct role in the degradation of the xenobiotic. Three examples suffice as illustration: 1. Enrichment in a two-stage chemostat with parathion (O,O-diethyl-O-(4-nitrophenyl)phosphorothioate) as the sole source of carbon and sulfur resulted in a community that was stable for several years (Daughton and Hsieh 1977). It should be noted that, on account of the toxicity of parathion to the culture, only low substrate concentrations could be used, and this methodology is ideally adapted to such situations. Degradation was accomplished by two organisms, Pseudomonas stutzeri and P. aeruginosa, whereas the third organism in the stable community had no defined function; P. stutzeri functioned only in ester hydrolysis, which is the first step in the degradation of parathion. This was the first demonstration of degradation of parathion by a metabolically defined microbial consortium, though degradation by a culture consisting of nine organisms had already been demonstrated (Munnecke and Hsieh 1976). 2. Chemostat enrichment was carried out with a mixture of linear alkylbenzene sulfonates as the sole sources of carbon and sulfur at a concentration of 10 mg/L, and resulted in the development of a four-component consortium (Jiménez et al. 1991). Three of the organisms were apparently necessary to accomplish this apparently straightforward degradation, though the isolation procedure that used a complex medium with glucose as carbon source is not entirely unequivocal. A similar situation arose with hexadecyltrimethylammonium chloride from which three strains that could grow with the substrate were obtained, again after streaking on yeast–glucose medium (Van Ginkel et al. 1992). 3. Chemostat enrichment with 2-chloropropionamide yielded a community of at least six organisms: one of these, a Mycoplana sp. carried out hydrolysis of the amide, while various other components used the resulting free acid for growth. An interesting observation was that after prolonged incubation at a dilution rate of 0.01/h, a single strain of Pseudomonas sp. capable of growth solely on 2-chloropropionamide as carbon source could be isolated (Reanney et al. 1983). There are some general conclusions that may be drawn from the results of these experiments: 1. Relatively simple reactions were involved in the degradations, and these can be expected to be available to single organisms. 2. The first stages for two of these degradations were straightforward hydrolytic reactions. 3. In all the examples, organisms with undefined metabolic functions were present and probably fulfilled an important role in providing complex organic substrates in the form of cell lysis products or nutritional requirements.

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Reaction sequences used for the degradation of xenobiotics in natural systems may therefore be more complex than might plausibly be predicted on the basis of studies with pure cultures using relatively high substrate concentrations. Attention should be drawn to experiments in which solutions of the substrate in a suitable mineral medium are percolated through soil that is used as the source of inoculum. This is one of the classical procedures of soil microbiology and has been exploited to advantage in studies on the degradation of a range of chlorinated contaminants in groundwater (Van der Meer et al. 1992). Apart from the fact that this mimics closely the natural situation and incorporates the features inherent in any enrichment methodology, this procedure offers a degree of flexibility that enables systematic exploration of the following: 1. The effect of varying redox conditions since by altering the gas-phase experiments can readily be carried out under aerobic, microaerophilic, or anaerobic conditions. 2. The effect of substrate concentration and the important issue of the existence or otherwise of threshold concentrations below which degradation is not effectively accomplished. 3. The influence of sorption/desorption on biodegradation, which has been discussed in a wider context in Chapter 4. Apart from its application to the specific problem of groundwater contamination, this procedure offers a potentially valuable procedure for simulating bioremediation of contaminated soils.

SIMULTANEOUS PRESENCE OF TWO SUBSTRATES Analogous to the fact that pure cultures of microorganisms seldom occur in natural ecosystems, it is very rare for a single organic substrate to exist in appreciable concentrations. The relevant microorganisms under natural situations are therefore exposed simultaneously to several compounds and this situation can be simulated in laboratory experiments. Although the term “cometabolism” has been used extensively, it has been applied to conflicting metabolic situations and the pragmatic term “concurrent metabolism” (Neilson et al. 1985) offers an attractive alternative when more than one substrate is present. Three examples are used to illustrate the application of this procedure to experiments in which the pathways for the biotransformation of different xenobiotics have been established: 1. Experiments have used cells with a metabolic capability that may plausibly be predicted as relevant to that of the xenobiotic. For example, although elective enrichment failed to yield organisms that were able to grow at the expense of dibenzo[1,4]dioxin, its metabolism could be studied using a strain of Pseudomonas sp. that was capable of growth with naphthalene (Klecka and Gibson 1979). Cells were grown with salicylate (1 g/L) in the presence of dibenzo[1,4]dioxin (0.5 g/L), and two metabolites of the latter were isolated: dibenzo[1,4]dioxin-cis-1,2-dihydrodiol and 2-hydroxydibenzo[1,4]dioxin. The former is consistent with the established dioxygenation of naphthalene and the role of salicylate as coordinate inducer of the relevant enzymes for conversion of naphthalene into salicylate. 2. An environmentally relevant situation may be simulated by the growth of an organism with a single substrate at a relatively high concentration with simultaneous exposure to a structurally unrelated xenobiotic present at a significantly lower concentration. A series of investigations has used growth substrates at concentrations of 200 mg/L and xenobiotic concentrations of 100 µg/L; it may reasonably be assumed that growth with compounds at the latter concentration was negligible in these experiments. For example, during growth of a stable anaerobic enrichment culture with 3,4,5-trimethoxybenzoate, 4,5,6-trichloroguaiacol was transformed into 3,4,5-trichlorocatechol that was further dechlorinated to 3,5-dichlorocatechol (Neilson et al. 1987).

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3. There has been considerable discussion on the mineralization of DDT, and in particular the biodegradation of the intermediate DDE. Cells of Pseudomonas acidovorans strain M3GY have been shown, however, during growth with biphenyl to degrade DDE with the fission of one ring and production of 4-chlorobenzoate (Hay and Focht 1998).

USE OF UNENRICHED CULTURES: UNDEFINED NATURAL CONSORTIA Laboratory experiments using natural consortia under defined conditions have particular value from several points of view. They are of direct environmental relevance, and their use minimizes the ambiguities in extrapolation from the results of studies with pure cultures. They provide valuable verification of the results of studies with pure cultures and make it possible to evaluate the extent to which the results of such studies may justifiably be extended to the natural environment. It should be appreciated, however, that in some cases the habitats from which the inoculum was taken might already have been exposed to xenobiotics so that “natural” enrichment (preexposure) could already have taken place. This has been discussed in Chapter 4. Extensive studies on the effect of substrate concentration and on the bioavailability of the substrate to the appropriate microorganisms have employed samples of natural lake water supplemented with suitable nutrients. There are few additional details that need to be added since the experimental methods are straightforward and present no particular difficulties. Considerable use has also been made of a comparable methodology to determine the fate of agrochemicals in the terrestrial environment. Because of the difficulty in obtaining pure cultures of anaerobic bacteria, use has been made of anaerobic sediment slurries in laboratory experiments. In some of these, although no enrichment was deliberately incorporated, experiments were carried out over long periods of time in the presence of contaminated sediments and adaptation of the natural flora to the xenobiotic during exposure in the laboratory might therefore have taken place. The design of these experiments may also inevitably result in interpretative difficulties. A few illustrations are provided: 1. Although the results of experiments on the dechlorination of pentachlorophenol (Bryant et al. 1991) enabled elucidation of the pathways to be elucidated, this study also revealed one of the limitations in the use of such procedures. Detailed interpretation of the kinetics of pentachlorophenol degradation using dichlorophenol-adapted cultures was equivocal due to carryover of phenol from the sediment slurries. 2. The biodegradation of acenaphthene and naphthalene under denitrifying conditions was examined in soil–water slurries (Mihelcic and Luthy 1988), though in this case only analyses for the concentrations of the initial substrates were carried out. In both of these examples, growth of the degradative organisms was supported at least partly by organic components of the soil and sediment, so that the physiological state of the cells could not be precisely defined. It would, therefore, be desirable to avoid ambiguity by using cultures in which the sediment is no longer present. This approach is illustrated by extensive investigations on the anaerobic dechlorination of chlorocatechols (Allard et al. 1991, 1994), and 2,3,5,6-tetrachlorobiphenyl (Cutter et al. 1998).

MICROCOSM EXPERIMENTS Microcosms are laboratory systems generally consisting of tanks such as fish aquaria containing natural sediment and water or soil. In those that have been most extensively evaluated for aquatic systems, continuous flow systems are used. In all of them, continuous measurement of 14CO2 evolved from 14C-labeled substrates may be incorporated, and recovery of both volatile and nonvolatile metabolites is possible so that a material balance may be constructed (Huckins et al. 1984). It should

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be pointed out that the term microcosm has also been used to cover much smaller-scale experiments that have been carried out in flasks under anaerobic conditions (Edwards et al. 1992), and to systems for evaluating the effect of toxicants on biota. Some examples are given to illustrate different facets of the application of microcosms to study various aspects of biodegradation: 1. Biodegradation of t-butylphenyl diphenyl phosphate was examined using sediments either from an uncontaminated site or from one having a history of chronic exposure to agricultural chemicals (Heitkamp et al. 1986). Mineralization was very much more extensive in the latter case, but was inhibited by substrate concentrations exceeding 0.1 mg/L. Low concentrations of diphenyl phosphate, 4-t-butylphenol, and phenol indicated the occurrence of esterase activity, while the recovery of triphenyl phosphate suggested dealkylation by an unestablished pathway. A comparable study of naphthalene biodegradation (Heitkamp et al. 1987) found more rapid degradation when sediments chronically exposed to petroleum hydrocarbons were used, and isolation of cis-1,2-dihydronaphthalene-1, 2-diol, 1- and 2-naphthol, salicylate, and catechol confirmed the pathway established for the degradation of naphthalene. The results of both investigations illustrate the potential for a more extensive application of the procedure and at the same time the significance of preexposure to the xenobiotic. 2. One of the key issues in bioremediation is the survival of the organisms deliberately introduced into the contaminated system. A microcosm prepared from a pristine ecosystem was inoculated with a strain of Mycobacterium sp. that had a wide capacity for degrading PAHs, and this organism was used to study the degradation of 2-methylnaphthalene, phenanthrene, pyrene, and benzo[a]pyrene (Heitkamp and Cerniglia 1989). The test strain survived in the system with or without exposure to PAHs, but the addition of organic nutrients was detrimental to its maintenance. Clearly, an almost unlimited range of parameters could be varied to enable a realistic evaluation of the effectiveness of bioremediation in natural circumstances. 3. Concern has been expressed on the potential hazards from discharge into the environment of organisms carrying catabolic genes on plasmids. A number of investigations (Jain et al. 1987; Fulthorpe and Wyndham 1989; Sobecky et al. 1992) have used a set of microcosms to determine the conditions needed to preserve the genotype and its stability. Once again the advantage of the technique is the ease of incorporating important variables that may be difficult to analyze in natural systems. 4. A sediment–water system was used to study the partition and the degradation of 14C-labeled 4-nitrophenol and 3,4-dichloroaniline (Heim et al. 1994). The results clearly illustrated the importance of water-to-sediment partitioning, and that a substantial fraction of the substrates existed in the form of nonextractable residues. 5. A study using resuspended river sediment (Marchesi et al. 1991) illustrated the important interdependence of substrate attachment to particulate matter and its biodegradability. Addition of sodium dodecyl sulfate that is degradable resulted in a relative increase in the number of particle-associated bacteria, whereas this was not observed with the nondegradable analogs such as sodium tetradecyl sulfate or sodium dodecane sulfonate. 6. A series of soil microcosms were used to study the biodegradation and bioavailability of pyrene during long-term incubation. The nonextractable fraction of 14C-labeled pyrene that had been introduced into pristine soil and incubated with or without the addition of azide was substantially greater in the latter (Guthrie and Pfaender 1998). It was also shown that microbial activity produced a number of unidentified polar metabolites that might plausibly be involved in the association. Truly field experiments on microbial reactions are extremely difficult to carry out, but a series of microcosm experiments on the substrates that may support anaerobic sulfate reduction quite closely

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approached this ideal situation (Parkes et al. 1989). This investigation used inhibition of sulfate reduction by molybdate to study the increase in the levels of a wide range of organic substrates endogenous in the sediments that were used. These substrates included a range of alkanoic acids and amino acids and, at the same time, considerably increased the range of organic substrates able to support sulfate reduction. Both in situ microcosms and laboratory systems were used to compare and evaluate first-order rates of degradation for a range of mixed substrates including aromatic hydrocarbons and phenolic compounds (Nielsen et al. 1996). The observed rates were comparable, although no systematic differences were observed with the exception of 2,6-dichlorophenol, which was not degraded in the laboratory system.

EXPERIMENTS IN MODELS OF NATURAL AQUATIC SYSTEMS It is extremely difficult to carry out field investigations in natural ecosystems with the rigor necessary to unravel metabolic intricacies, although such experiments have been successfully carried out in investigations aimed at determining the fate and persistence of agrochemicals in the terrestrial environment, and in the context of bioremediation. In general, however, simplified systems have been developed. These attempt to simulate critical segments of natural ecosystems in a clearly defined way. Outdoor model systems have been used and two examples are used to illustrate the kinds of data that can be assembled and the range of conclusions—and their limitations—that may be drawn from such experiments. Not only purely microbiological determinants of persistence may be revealed, but also important data on the distribution and fate of the xenobiotic may be acquired. 1. Studies in an artificial stream system were designed to provide confirmation in a field situation of the results from laboratory experiments that had demonstrated the biodegradability of pentachlorophenol. Pentachlorophenol was added continuously to the system for 88 d and its degradation followed (Pignatello et al. 1983, 1985). The results confirmed that pentachlorophenol was indeed degraded by the natural populations of microorganisms and, in addition, drew attention to the significance of both sediments and surfaces in the partitioning of pentachlorophenol between the phases within the system. 2. 4,5,6-Trichloroguaiacol was added continuously during several months to mesocosm systems simulating the Baltic Sea littoral zone. Samples of water, sediment, and biota including algae were removed periodically for analysis both of the original substance and of metabolites identified previously in extensive laboratory experiments (Neilson et al. 1989). A complex of metabolic transformations of 4,5,6-trichloroguaiacol was identified, including O-methylation to 3,4,5-trichloroveratrole, O-demethylation to 3,4,5-trichlorocatechol, and partial dechlorination to a dichlorocatechol, and these metabolites were distributed among the various matrices in the system. Of particular significance was the fact that a material balance unequivocally demonstrated the role of the sediment phase as a sink for both the original substrate and the metabolites, so that a number of interrelated factors determined the fate of the initial substrate. 3. Mesocosms placed in shallow Finnish lakes were used to evaluate changes brought about by extended incubation of biologically treated bleachery effluent from mills that used chloride dioxide. The mesocosms had a volume of ca. 2 m3 and were constructed of translucent polyethere or black polyethene to simulate dark reactions. The experiments were carried out at ambient temperatures throughout the year, and sum parameters were used to trace the fate of the organically bound chlorine. In view of previous studies on the molecular mass distribution of effluents (Jokela and Salkinoja-Salonen 1992), this was measured as an additional marker. Important features were that (a) sedimentation occurred exclusively within the water mass within the mesocosm, (b) the atmospheric input could be estimated

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from control mesocosms, and (c) the microbial flora included both indigenous organisms in lake water and those carried over from the treatment process (Saski et al. 1996a,b). There were a number of important general conclusions: • The atmospheric input of organically bound halogen was negligible compared with that of the effluent. • There was >50% loss of organohalogen in the water phase and this included compounds with masses of both <500 and >500. • The tetrahydrofuran-extractable organic chlorine in the de novo sediment had a molecular mass (average 1400) that was much higher than that from the water-phase extracts (average 360). There was no evidence for selective sorption of higher-molecular-mass components although transformation of lower-molecular-mass components to highermolecular-mass hydrophobic compounds cannot be excluded. 4. A mesocosm was constructed to study the dynamics of populations of bacteria in an aquifer contaminated with BTEX (principally benzene). Water from the site was contained in permeable tubes placed within perforated polypropylene cylinders at sites within the aquifer. Periodic samples were removed and analyzed to characterize the bacterial community using denaturing gel electrophoresis, analysis of PCR-amplified 16S rRNA genes, and BTEX degradative genes (Hendricks et al. 2005). The results showed that a stable bacterial community developed downstream after access to bacteria from the contaminated site, and that the community was characterized by genotypes similar to xylM/xylE1.

EVALUATION OF DEGRADATION USING METABOLITES Comparable experiments in natural aquatic ecosystems are generally difficult to design (Madsen 1991), although some examples of what may be accomplished are given as illustration, and are applicable when there is sufficient knowledge about the degradative pathways of the xenobiotic: 1. Analysis of chlorobenzoates in sediments, which had been contaminated with PCBs, was used to demonstrate that the lower PCB congeners that had initially been produced by anaerobic dechlorination were subsequently degraded under aerobic conditions. The chlorobenzoates were transient metabolites and their concentrations were extremely low since bacteria that could successfully degrade them were present in the sediment samples (Flanagan and May 1993). 2. The bacterial aerobic degradation of pyrene is initiated by the formation of cis-pyrene4,5-dihydrodiol. Analysis for this metabolite was used to demonstrate the biodegradability of pyrene in an environment in which there was continuous input of the substrate, when it was not possible to use any diminution in its concentration as evidence for biodegradation (Li et al. 1996). The corresponding metabolite from naphthalene—cis-naphthalene1,2-dihydrodiol—has been used to demonstrate biodegradation of naphthalene both in site-derived enrichment cultures and in leachate from the contaminated site (Wilson and Madsen 1996). 3. It has been shown that pure cultures of bacteria under anaerobic denitrifying conditions may produce benzylsuccinate as a metabolite of toluene (Evans et al. 1992; Migaud et al. 1996; Beller et al. 1996). Demonstration of this and the corresponding methylbenzyl succinates from xylenes has been used to demonstrate metabolism of toluene and xylene in an anaerobic aquifer (Beller et al. 1995, 2002). However, care must be exercised in the interpretation of results that show the presence of putative metabolites. An illustrative example is provided by a study of the biodegradation

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of a range of PAHs in compost-amended, unamended, and sterilized soil (Wischmann and Steinhart 1997): a. Neither dihydrodiols formed by bacterial dioxygenation nor phenols from fungal monooxygenation followed by rearrangement or hydrolysis and elimination were found. b. In contrast, plausible oxidation products of anthracene, acenaphthylene, fluorene, and benz[a]anthracene—anthracene-9.10-quinone, acenaphthene-9,10-dione, fluorene-9-one, and benz[a]anthracene-7,12-quinone—were found transiently in compost-amended soil. It was shown, however, that these were formed even in sterile controls by undetermined abiotic reactions. These results clearly illustrate the care that must be exercised in interpreting the occurrence of PAH oxidation products as evidence of biodegradation. On account of the potential health hazard, the application of radiolabeled substrates, which have been the cornerstone of metabolic experiments, is not generally acceptable. Examples of experiments using stable isotopes have, however, been carried out to determine biodegradation under field conditions and to establish the source of contaminants. The application of stable isotope fractionation is discussed in Chapter 6, Part 1, and a single illustration of the use of deuterium-labeled substrates is given here. Fully deuterated benzene, toluene, 1,4-dimethylbenzene, and naphthalene were used to determine their dissemination and biodegradation in an aquifer plume with bromide as an inert marker (Thierrin et al. 1995). Analysis of samples was readily accomplished by GC-MS, and after taking into account loss by sorption and dispersion, the half-lives of the substrates were calculated. At the oxic upper surface of the plume, rapid degradation occurred and continued at a slower rate into the anoxic zone. Although benzene was the most persistent substrate, there is evidence that it is degradable under anaerobic methanogenic, sulfate-reducing, Fe(III)-reducing (Kazumi et al. 1997), and nitrate-reducing (Burland and Edwards 1999) conditions.

EXPERIMENTAL PROBLEMS: WATER SOLUBILITY, VOLATILITY, SAMPLING, AND ASSOCIATION OF THE SUBSTRATE WITH MICROBIAL CELLS Although these issues have already been briefly noted, they deserve a few additional comments. For freely water-soluble substrates that have low volatility, there are few difficulties in carrying out the appropriate experiments described above. There is, however, increasing interest in xenobiotics such as polycyclic aromatic hydrocarbons (PAHs) and highly chlorinated compounds including, for example, PCBs, which have only low water solubility. In addition, attention has been focused on volatile chlorinated aliphatic compounds such as the chloroethenes, dichloromethane, and carbon tetrachloride. All of these substrates present experimental difficulties of greater or lesser severity. 1. Whereas suspensions of poorly water-soluble substrates can be used for experiments on the identification of metabolites, these methods are not suitable for kinetic experiments that necessitate the quantification of substrate concentrations. In such cases, the whole sample must be sacrificed at each sampling time, and care must be taken to ensure that substrate concentrations in each incubation vessel are as far as possible equal. In addition, the whole sample must be extracted for analysis since representative aliquots cannot be removed. Experiments are initiated by preparing solutions of the substrates in suitable solvents such as acetone, diethyl ether, methanol or ethyl acetate, sterilizing the solutions by filtration, and dispensing appropriate volumes into each incubation vessel. The solvent is then removed in a stream of sterile N2 and the cell suspension added. As an alternative, solutions in any of these solvents or in others that are much less volatile, such as dimethyl sulfoxide and dimethyl formamide, have been added directly to media after sterilization by filtration. However, since the remaining concentration of the solvents may not

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be negligible, care must be taken to ensure that these neither toxic nor compromise the results of the metabolic experiments. When the test substrate is a solid, it may be preferable to prepare saturated solutions in the basal medium, remove undissolved substrate by filtration through glass-fiber filters, and sterilize the solution by filtration before dispensing and adding cell suspensions. An attractive procedure has used PAHs sorbed onto sheets of hydrophobic membranes and incubated with samples of contaminated soils (Bastiaens et al. 2000). After rinsing the sheets to remove particulate material, the membranes were placed on agar mineral medium with or without addition of PAHs. After incubation, cultures were then made from the cell mass that had developed. 2. Investigation of substrates with appreciable volatility—and many organic compounds including solids have significant vapor pressures at ambient temperatures—presents a greater experimental challenge, especially if experiments are to be conducted over any length of time. Incubation vessels such as tubes or bottles may be closed with rubber stoppers or with teflon-lined crimp caps fitted with rubber seals, and these are particularly convenient for withdrawing samples using syringes. Even with Teflon seals, however, these cap inserts may be permeable to compounds with appreciable vapor pressure, and sorption of the test substrate may also occur to a significant extent. Both of these factors result in controls that display undesirable diminution in the concentration of the test substrate. Good illustrative examples are provided by the results of a study with endosulfans and related compounds (Guerin and Kennedy 1992), and on the metabolism of 4,4′dichlorobiphenyl by Phanerochaete chrysosporium (Dietrich et al. 1995). These results support the importance even in laboratory experiments of taking into account gas/liquid partitioning. Completely sealed glass ampoules may be used for less volatile compounds—though clearly not for highly volatile compounds—and subsampling cannot be carried out. For aerobic organisms, one serious problem with all these closed systems is that of oxygen limitation. The volume of the vessels should therefore be very much greater than that of the liquid phase, and conclusions from the experiments should recognize that microaerophilic conditions will almost certainly prevail during prolonged incubation. Such limitations are clearly not relevant for anaerobic organisms. 3. It is well established that associations occur between xenobiotics and biota, and these may introduce ambiguities into the interpretation of the results of metabolic experiments using microorganisms. Three illustrative examples are given for association of xenobiotics with microorganisms: a. Chloroguaiacols with Gram-positive bacteria (Allard et al. 1985) b. PCB congeners with the hyphae of Phanerochaete chrysosporium (Dietrich et al. 1995) c. 2,4,6-Trichlorophenol with Bacillus subtilis (Daughney and Fein 1998) Analytical procedures should, therefore, be designed to take into account such associations. If they are not quantitatively evaluated or eliminated, the results of such experiments may be seriously compromised.

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6

Elucidation of Metabolic Pathways

INTRODUCTION Two sections deal briefly with procedures for investigating the pathways used in biodegradation and biotransformation. They cover briefly the application of isotopes, and of nondestructive methods that include NMR, EPR, and x-ray analysis. They should be viewed in the wider context of procedures for evaluating the effectiveness of bioremediation that are covered in Chapter 13 and their application in Chapter 14.

PART 1 APPLICATION OF NATURAL AND SYNTHETIC ISOTOPES The application of substrates isotopically labeled in specific positions makes it possible to follow the fate of individual atoms during the microbial degradation of xenobiotics. Under optimal conditions, both the kinetics of the degradation, and the formation of metabolites may be followed— ideally when samples of the labeled metabolites are available. Many of the classical studies on the microbial metabolism of carbohydrates, carboxylic acids, and amino acids used radioactive 14 C-labeled substrates and specific chemical degradation of the metabolites to determine the position of the label. The method is indeed obligatory for distinguishing between degradative pathways when the same products are produced from the substrate by different pathways. A suitable example is provided by the β-methylaspartate and hydroxyglutarate pathways for fermentations of glutamate, both of which produce butyrate but which could clearly be distinguished by the use of [4-14C]glutamate (Buckel and Barker 1974).

CARBON (14C AND 13C) Conventional use has been made of the radioisotope 14C, and details need hardly be given here. Illustrative examples include the elucidation of pathways for the anaerobic degradation of amino acids (Chapter 7, Part 1) and purines (Chapter 10, Part 1). Some applications have used 13C with high-resolution Fourier transform NMR in whole-cell suspensions, and this is equally applicable to molecules containing the natural 19F or the synthetic 31P nuclei. As noted later, major advances in NMR have made it possible to use natural levels of 13C. Illustrative examples of the application of 13C-labeled substrates to biodegradation include the following: 1. The degradation of 13CCl4 by Pseudomonas sp. strain KC involved formation of intermediate COCl2 that was trapped as a HEPES complex, and by reaction with cysteine (Lewis and Crawford 1995). Further details of the pathway that is mediated by the metabolite pyridine-dithiocarboxylic acid have been elucidated (Lewis et al. 2001). 2. 13C[bicarbonate] and NMR were used to demonstrate that the first product in the metabolism of propene epoxide is acetoacetate, which is then reduced to β-hydroxybutyrate (Allen and Ensign 1996). 3. 13C[bicarbonate] and mass spectrometry were used to demonstrate the formation of carboxylic acids during the sulfidogenic mineralization of naphthalene and phenanthrene (Zhang and Young 1997).

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4. 9[13C]-anthracene was used to study its degradation in soil, and the formation of labeled metabolites that could be released only after alkaline hydrolysis (Richnow et al. 1998). It was possible to construct a carbon balance during the 599-d incubation, and to distinguish metabolically formed phthalate from indigenous phthalate in the soil. 5. Traps with Bio-Sep beads amended with [13C6]-benzene and [13C]-toluene were used to assess biodegradation in an aquifer (Geyer et al. 2005). Beads were lyophilized after exposure, lipids were extracted with chloroform–methanol, and the fatty acids and δ13C values analyzed. High enrichment of 13C was observed in several fatty acids, which showed that the label from the substrates had been incorporated. In addition, there were differences in the abundance of the fatty acids in beads amended with benzene or toluene that suggested the existence of different microbial degradative populations. 6. The degradation of 13C6-benzene was studied in anaerobic enrichment cultures when phenol, benzoate, and toluene were detected, and the kinetics of their formation studied (Ulrich et al. 2005).

SULFUR (35S) AND CHLORINE (36Cl) Although quite extensive use of 35S has been made in studies on the degradation of alkyl sulfonates (Hales et al. 1986), 36Cl has achieved only limited application on account of technical difficulties resulting from the low specific activities and the synthetic inaccessibility of appropriately labeled substrates. One of the few examples of its application to the degradation of xenobiotics is provided by a study of the anaerobic dechlorination of hexachlorocyclohexane isomers (Jagnow et al. 1977), the results of which are discussed in Chapter 7, Part 3.

HYDROGEN (2H) AND OXYGEN (18O) Although the radioactive isotope 3H has been extensively used for studies on the uptake of xenobiotics into whole cells, the intrusion of exchange reactions and the large isotope effect renders this isotope rather less straightforward for metabolic studies. Both deuterium 2H-labeled substrates, and oxygen 18O2 and 18OH2 have, however, been extensively used in metabolic studies, since essentially pure labeled compounds are readily available and mass spectrometer facilities have become an essential part of structural determination. Illustrative examples of the use of 2H in different applications include the following: 1. Deuterium labeling has been invaluable in studying rearrangements involving protons. For example, it has been used to reveal the operation of the NIH shift during metabolism of [4-2H]ethylbenzene by the monooxygenase system from Methylococcus capsulatus (Dalton et al. 1981), of [2,2′,3,3′,5,5′,6,6′-2H]biphenyl by Cunninghamella echinulata (Smith et al. 1981), and of [1-2H]naphthalene and [2-2H]naphthalene by Oscillatoria sp. (Narro et al. 1992). 2. The conversion of long-chain alkanoate CoA esters into the alkenoate CoA esters by acyl-CoA oxidase involves an anti elimination reaction. The stereochemistry of the reaction in Candida lipolytica was established using stearoyl-CoA–labeled with 2H at the 2(R)-, 3(R)-, and 3(S)-positions (Kawaguchi et al. 1980). 3. The use of 2H-labeled substrates has been used to determine details of the dehydrogenation of cis-dihydrodiols produced by dioxygenases from aromatic substrates (Morawski et al. 1997), and it was possible to demonstrate the specificity of hydrogen transfer from the dihydrodiol substrates to NAD. 4. The fumarate pathway for the metabolism of alkanes under sulfate-reducing conditions was examined using fully deuterated hexadecane (Callaghan et al. 2006). Identification of the metabolites including labeled methylpentadecylsuccinate and 4-methyloctadecanoate

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was used to determine details of the pathway that involved a rearrangement during the initial reaction. The enantiomerization of phenoxyalkanoic acids containing a chiral side chain has been studied in soil using 2H2O (Buser and Müller 1997). It was shown that there was an equilibrium between the R- and S- enantiomers of 2-(4-chloro-2-methylphenoxy)propionic acid (MCPP) and 2-(2,4-dichlorophenoxy)propionic acid (DCPP) with an equilibrium constant favoring the herbicidally active R-enantiomer. The exchange reactions proceeded with both retention and inversion of configuration at the chiral sites. 2 H-labeled substrates have been used to determine the dissipation and degradation of aromatic hydrocarbons in a contaminated aquifer plume (Thierrin et al. 1995). Its application was particularly appropriate since the site was already contaminated with the substrates. With suitable precautions, this procedure seems capable of extension to determining the presence—though not the complete structure—of metabolites, provided that the possibility of exchange reactions were taken into account. The fate of toluene and o-xylene in an aquifer contaminated with BTEX was examined by injecting toluene-d8 and o-xylene-d10 followed by quantification of the label in benzyl succinate and 2-methylbenzene succinate (Reusser et al. 2002) that are established metabolites on the anaerobic pathway for the degradation of toluene and o-xylene. Styrene-degrading bacteria from full-scale and experimental biofilters were exposed to [2H8]styrene, and analysis of fatty acids was used to distinguish the bacterial flora of the two systems (Alexandrino et al. 2001). Biodegradation of toluene was evaluated at a site that was contaminated with benzene and toluene from the earlier production of benzene. Anaerobic degradation was supported with sulfate as the electron acceptor, and ring-deuterated (d5) and fully deuterated (d8) toluene were injected into the anoxic zone. Biodegradation was assessed by several procedures such as (a) isotope fractionation using the Raleigh equation combined with laboratorydetermined isotope fractionation and (b) analysis of changes in the concentration of the deuterated toluene relative to bromide that was used as a marker. Both procedures yielded comparable results, and either of them was able to provide a record of toluene biodegradation. Degradation was confirmed by demonstration of benzylsuccinate-d 5 and an initial increase in the ratio of deuterium to hydrogen in the groundwater (Fischer et al. 2006).

Although application of 18O has been less frequent, it has been used effectively to determine the source of oxygen and the number of oxygen atoms incorporated during metabolism of xenobiotics under both aerobic and anaerobic conditions. Some typical examples are given below: 1. In a classical study, it was shown that during bacterial oxidation of benzene to catechol both atoms of oxygen came from 18O2 (Gibson et al. 1970). This initiated the appreciation of the role of dioxygenases in the degradation of aromatic xenobiotics, and many examples are given in Chapter 8, Parts 1 and 2. 2. During the biodegradation of 2,4-dinitrotoluene by a strain of Pseudomonas sp., two atoms of oxygen were incorporated from 18O2 during the formation of 4-methyl-5-nitrocatechol by dioxygenation with loss of nitrite (Spanggord et al. 1991). 3. The degradation of nicotine has been examined extensively in Arthrobacter nicotinovorans (oxydans). In strain P34, the first metabolite was 6-hydroxynicotine, and experiments with 18 O2 and H218O showed that the oxygen in the hydroxyl group was derived from H2O (Hochstein and Dalton 1965). 4. Experiments with H218O using Pseudomonas putida strain 86 (Bauder et al. 1990) showed that the oxygen incorporated into quinol-2-one originates from water.

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5. During the degradation of 2-chloroacetophenone by a strain of Alcaligenes sp., one atom of 18O2 was incorporated into 2-chlorophenol formed from the 2-chlorophenyl acetate that was initially formed by Baeyer–Villiger monooxygenation (Higson and Focht 1990). 6. Benzene and toluene were anaerobically hydroxylated to phenol and 4-hydroxytoluene, and experiments with H218O showed that the oxygen atoms had originated from water (Vogel and Grbic-Galic 1986). 7. The mechanism of the cytochrome P45014DM conversion of a C14 methyl sterol into formate and the ∆14,15 was studied using 18O2, and it could be shown that the hydroxyl oxygen atom in the formate that was produced contained one atom of 18O (Shyadehi et al. 1996). In experiments involving the use of 18O2 and H218O, care should be taken to exclude chemical exchange reactions involving potentially labile C–H or C–O bonds, since these reactions could seriously compromise the conclusions. A good illustration of the pitfalls in such investigations is shown by a study on the dechlorination of pentachlorophenol by the dehalogenase from a strain of Arthrobacter sp. The initial reaction in the degradation of pentachlorophenol is mediated by a pentachlorophenol dehalogenase that produces tetrachloro-1,4-dihydroxybenzene. Experiments using the enzyme showed that 18O is incorporated into this metabolite only after incubation with H218O and not with 18O2. It has been shown that the labeling occurs as a result of exchange between the initially formed unlabeled metabolite and H218O. An unambiguous elucidation of the mechanism of the reaction was not therefore possible, since even if 18O had been incorporated during the reaction with 18O2, exchange with the excess H216O in the medium would have yielded an unlabeled product (Schenk et al. 1990).

OTHER ISOTOPES Studies of biodegradation have made only limited use of metal isotopes: 1. The aerobic degradation of the 2-carboxylates of furan, pyrrole, and thiophene is initiated by hydroxylation before fission of the rings. Although details of the enzymes are limited, it was suggested on the basis of tungstate inhibition and 185W [tungstate] labeling that the degradation of 2-furoyl-coenzyme A involves a molybdenum-dependent dehydrogenase (Koenig and Andreesen 1990). 2. Use of Mössbauer spectra generally requires the addition of 57Fe to the growth medium. This procedure has received extensive application to studying the environment of Fe in enzymes, and illustrative examples include the following: a. EPR analysis and Mössbauer spectra of the enoate reductase in Clostridium tyrobutyricum revealed the presence of a semiquinone and an [4Fe–4S] cluster (Caldeira et al. 1996). b. Mössbauer spectra of the sulfate reductase of Bacillus subtilis showed the presence of a [4Fe–4S] cluster in the isolated enzyme, and that this was degraded to a [2Fe–2S] cluster. It was suggested that the interconversion of these clusters functioned in suppressing formation of sulfite under aerobic conditions (Berndt et al. 2004). 3. The anaerobic degradation of nicotinate by Clostridium barkeri is initiated by nicotinate hydroxylase that introduces oxygen from water at C6. The active enzyme contains Se, and it was shown that in cultures grown with 77Se the signal of Mo(V) in the EPR spectrum was split without affecting the Fe signals of the iron–sulfur cluster (Gladyshev et al. 1994). The selenium was therefore coordinated with the molybdopterin cofactor.

ISOTOPE EFFECTS AND STABLE ISOTOPE FRACTIONATION These procedures have had two principal applications: analysis of degradation and transformation in natural populations, and the determination of degradation pathways. Illustrative examples are

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given in Chapter 13 and only a brief outline of experimental aspects is attempted. Depending on the objective, results from these experiments have been reported in a number of ways, and in comparing results it is important to take this into consideration. This aspect is therefore discussed here. Elements such as C, N, O, S, and Cl that are components of many organic compounds exist naturally as mixtures of stable isotopes. The ratios of these in a compound reflect the different rates of reaction at isotopically labeled positions, and therefore reflect the fractionation—biotic or abiotic—by which it was synthesized or to which the compound has been subjected. Techniques have been developed whereby the ratios 13C/12C (δ 13C), 15N/14N (δ 15N), 18O/16O (δ 18O), 34S/32S (δ 34S), and 37Cl/35Cl (δ 37Cl) can be accurately measured by mass spectrometry. The differences are expressed as per mille (‰) deviations with respect to a standard: δ = 1000 × (ratio for sample/ratio for standard) – 1 Standard samples are used for calibration, and procedures for C and O have been given in detail by Craig (1957), and include details of the standards, the appropriate correction factors that should be applied, and mass spectrometric techniques. The standard for carbon is the Vienna Pee Dee Belemnite (VPDM), and for oxygen Vienna Standard Mean Ocean Water (VMOW). Secondary standards have included graphite and atmospheric oxygen that have been related to the primary standards. Atmospheric air has been used for nitrogen and ocean water chloride for chlorine. These values may differ slightly from those derived from the relative isotope contributions given in compilations of the elements. Determination of δ 13C has been used extensively in geochronology, and δ 18O in glaciology and climatology, and these techniques have also been used in determining biosynthetic pathways of, for example, lipids, and the various processes whereby CO2 is incorporated into and disseminated in biota (Hayes 1993). They have been used increasingly in environmental research. Some examples are given to illustrate the potential of this methodology, using analysis of individual compounds to establish their sources, or alterations during transformation or dissemination. Extensive applications have been made to the biodegradation of chlorinated ethenes, monocyclic aromatic hydrocarbons, polychlorinated biphenyls, and methyl tert-butyl methyl ether. They have been directed to anaerobic reactions since these conditions are widespread in contaminated groundwater systems (Chapter 14, Part 7). The results have been reported in a number of ways, and valuable comments on pitfalls that may be encountered when heavy isotopes are used have been given (Hunkeler 2002). Results have been expressed in a number of ways. In the Raleigh model that has been extensively used, the fractionation factor α is given by R/R0 = f (α – 1) when the fraction of remaining substrate is f and where R is the isotopic composition of the substrate during degradation and R0 is the initial value. The enrichment factor ε where ε = 1000 (α – 1) has also been used. There are certain conditions that must be fulfilled for the Raleigh model to be applicable: 1. The concentrations of the heavy isotope must be low, which is true for experiments using natural levels of abundance, or the reaction must be near the initial starting point. 2. The model is strictly applicable only when applied to a single-step reaction, and not to a sequence of reactions. Kinetic isotope effects have also been given in terms of ε; for reactions at a single carbon atom 12 13 k/ k = 1/(1 + ε/1000) where 12 k and 13k are rates for 12C and 13C substrates. In the application of the Raleigh equation to contaminated sites, complications due to the heterogeneity of the system have been addressed (Abe and Hunkeler 2006). Detailed analysis has revealed systematic problems that could result in underestimation of first-order rates, and procedures for evaluating these were provided.

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Experimental Procedures The analyte must be converted into a volatile compound suitable for mass-spectrometric analysis. Procedures for C, N, and O follow those developed for conventional organic microanalysis— oxidation of organic C to CO2, reduction of organic N to N2, and conversion of O2 into CO or CO2. In most procedures, cryogenic purification of the products is carried out before mass spectrometry, and both off-line and on-line procedures have been developed. Carbon. Determination of δ 13C has been carried out by GC separation of the analytes followed by high-temperature oxidation. Although oxidation with CuO at 850°C has been used, use of NiO at 1050°C supplemented by O2 is preferable, and it is important to remove H2O before isotopic analysis (Merritt et al. 1995). Nitrogen. For δ 15N, the sample is heated at 850°C with an excess of Cu (to reduce NOx to N2) and CuO, followed by cryogenic purification of the N2 (Kucklick et al. 1996). Oxygen. For determination of δ 18O, the compound is pyrolyzed with graphite embedded with platinum wire at 800°C (Aggarwal et al. 1997) or glassy carbon at 1080°C (Saurer et al. 1998) for conversion to CO2 and CO, respectively. Chlorine. For the determination of δ 37Cl in organic compounds, the following methods have been used: a. Conversion to CuCl and then into CH3Cl by reaction with CH3I at 300°C (Holt et al. 1997) b. Reduction to chloride with sodium biphenyl followed by conversion to CsCl and determination by thermal ionization mass spectrometry (Numata et al. 2002) c. Conversion to AgCl followed by ionization by xenon atoms and determination of isotopes using an FAB-isotope ratio mass spectrometer (Westaway et al. 1998)

REFERENCES Abe Y, D Hunkeler (2006) Does the Raleigh equation apply to evaluate field isotope data in contaminant hydrogeology? Environ Sci Technol 40: 1588–1596. Aggarwal PK, ME Fuller, MM Gurgas, JF Manning, MA Dillon (1997) Use of stable oxygen and carbon isotope analyses for monitoring the pathways and rates of intrinsic and enhanced in situ biodegradation. Environ Sci Technol 31: 590–596. Alexandrino M, C Knief, A Lipski (2001) Stable-isotope based labeling of styrene-degrading microorganisms in biofilters. Appl Environ Microbiol 67:4796–4804. Allen JR, SA Ensign (1996) Carboxylation of epoxides to β-keto acids in cell extracts of Xanthobacter strain Py2. J Bacteriol 178: 1469–1472. Bauder R, B Tshisuaka, F Lingens (1990) Microbial metabolism of quinoline and related compounds VII Quinoline oxidoreductase from Pseudomonas putida: a molybdenum-containing enzyme. Biol Chem Hoppe-Seyler 371: 1137–1144. Berndt C, CH Lillig, M Wollenberg, E Bill, MC Mansilla, D de Mendoza, A Seidler, JD Schwenn (2004) Characterization and reconstitution of a 4Fe–4S adenyl sulfate/phosphoadenyl sulfate reductase from Bacillus subtilis. J Biol Chem 279: 7850–7855. Buckel W, HA Barker (1974) Two pathways of glutamate fermentation by anaerobic bacteria. J Bacteriol 117: 1248–1260. Buser H-R, MD Müller (1997) Conversion reactions of various phenoxyalkanoic acid herbicides in soil 2 Elucidation of the enantiomerization process of chiral phenoxy acids from incubation in a D2O/soil system. Environ Sci Technol 31: 1960–1967. Caldeira J, R Feicht, H White, M Teixeira, JJG Mourat, H Simon, I Moura (1996) EPR and Mössbauer spectroscopic studies on enoate reductase. J Biol Chem 271: 18743–18748. Callaghan AV, LM Gieg, KG Kropp, JM Suflita, LY Young (2006) Comparison of mechanisms of alkane metabolism under sulfate-reducing conditions among two bacterial isolates and a bacterial consortium. Appl Environ Microbiol 72: 4274–4282.

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PART 2 NONDESTRUCTIVE METHODS FOR STRUCTURE DETERMINATION APPLICATION OF NMR NMR in its various forms has been increasingly used. This is largely the result of technical developments that are briefly summarized in the following paragraphs: 1. During the years since its introduction, there have been substantial increases in the magnetic field strengths that are attainable. Greater sensitivity and resolution are thereby achieved with modern instruments, so that sample size has become less significant—provided that the sample is of adequate purity. Application to naturally occurring levels of 13C has become routine, and examination of compounds with 15N, 17O, and 19F has been reported. NMR has been increasingly used for in vivo investigations of microbial degradation, since samples containing cell suspensions and the 13C-labeled substrates can be examined directly, and the kinetics of the appropriate resonance signals monitored. A 13C NMR method has been developed to determine the transformation and degradation of chloroacetonitrile and related compounds in various samples of soil and water (Castro et al. 1996). This would be amenable to extension to a much wider range of compounds and reactions. 2. The size of sample required has been reduced by a number of technical developments including micro inverse probes and micro cells (references in Martin et al. 1998), and has been reduced even further using a newly developed 1.7-mm submicro inverse-detection gradient probe (Martin et al. 1998). The combined use of inverse detection probes with solenoid microcoils has also been developed to reduce sample volumes for 13C NMR (Subramanian and Webb 1998). 3. Multidimensional spectra as well as techniques including DEPT (distortionless enhancement by polarization transfer), COSY (correlated spectroscopy), and ROESY (rotating-frame overhauser enhancement spectroscopy) have been increasingly used.

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4. Mass spectra are often carried out after high-performance liquid-chromatography (HPLC) separation of components, and when soft ionization techniques are used, they provide the molecular mass, but relatively little structural information. The combined application of an interface to NMR and to MS would provide a powerful structural tool, and for some mass-spectrometry methods that provide limited structural information this is a particularly attractive combination. Identification by NMR provides detailed structural information, and a system for the analysis of peptides involves initial HPLC followed by splitting of the fractions for subsequent MS and NMR analysis in a flow-through system (Holt et al. 1997). Application to environmental samples after fractionation is attractive, and a configuration has been developed for combining capillary zone electrophoresis, capillary HPLC, and capillary electrochromatography with NMR including use in stopped-flow experiments (Pusecker et al. 1998). Direct coupling to LC systems has been achieved, and the application to whole cells has been possible in a number of applications. Both the continuous- and stopped-flow modes have proved valuable, and a combination of continuous flow HPLC interfaced with 1H NMR and HPLC interfaced to thermospray mass spectrometry has been used to identify the products from the photolysis of 2,4,6-trinitrotoluene (Godejohann et al. 1998). The application of NMR analysis coupled to HPLC and supercritical fluid procedures has been made possible by a radical change in the design of the NMR instrument (Albert 1995), and further technical developments are likely to offer a wider range of applications. The procedure has been valuable for identifying the urinary metabolites of pharmaceuticals, which has used both continuous- and stopped-flow methods (Sidelmann et al. 1997). There are many advantages of NMR including the fact that it is nondestructive. It may be carried out not only in extracts after purification by conventional means, but also directly in culture supernatants, and even during their synthesis in the spectrometer tubes. These may be equipped with gas inlets, which make it possible to carry out studies under virtually any metabolic conditions. In addition, there are no experimental restrictions in handling radioactive material, though for studies of carbon metabolism access to 13C-labeled substrates is advantageous. Increasingly, however, many of these are equally accessible as their 14C analogues. However, the relatively low sensitivity of the method may preclude identification of metabolites that are formed transiently only in low concentration. Hydrogen 1H This has received widespread application, and a single example has been chosen as representative. The pathway for the degradation of morpholine by Mycobacterium aurum MO1 and Mycobacterium strain RP1 was examined using whole cells, and this confirmed its identity to the one that had been proposed earlier for M. chelonae (Combourieu et al. 1998; Poupin et al. 1998). Carbon 13C Interpretative difficulties may arise from the inherent design of NMR experiments that may necessitate the use of high substrate concentrations due to the relatively low sensitivity of the procedure. A single illustrative example is given of the occurrence of the artifacts that may be encountered. During a study of the metabolism of mandelate by Pseudomonas putida (Halpin et al. 1981), benzyl alcohol was unexpectedly identified when experiments were carried out at a high substrate concentration (50 mM). This was, however, subsequently shown to be due to the action of a nonspecific alcohol dehydrogenase under the anaerobic conditions prevailing at the high substrate concentration that was used for the identification of the metabolites (Collins and Hegeman 1984).

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C NMR using whole cells has been applied to the study of a number of relatively straightforward metabolic reactions involving small molecules that include the following: 1. Glycolysis in Saccharomyces cerevisiae (den Hollander et al. 1986) 2. Reduction of dimethyl sulfoxide and trimethylamine-N-oxide by Rhodobacter capsulatus (Rhodopseudomonas capsulata) (King et al. 1987) 3. Metabolism of acetate and methanol in Pseudomonas sp. (Narbad et al. 1989) 4. Nicotinate and pyridine nucleotide metabolism in Escherichia coli and Saccharomyces cerevisiae (Unkefer and London 1984) 5. Combined use of 1H and 13C NMR in a powerful combination to deduce the pathways of degradation of chloroethene (vinyl chloride) (Castro et al. 1992a,b) 6. Incorporation of 13CO2 into poly-β-hydroxybutyrate by a strain of Xanthomonas sp. that metabolizes propene or its oxide (Small and Ensign 1995; Allen and Ensign 1996) 7. Degradation of acetonitrile by Methylosinus trichosporium (Castro et al. 1996) 8. Metabolism of 13CCl4 by Pseudomonas sp. strain KC (Lewis and Crawford 1995) Elucidation of more complex metabolic pathways includes the following: 1. Detection of glycolate and 2-(2-aminoethoxy)acetate as intermediates in the degradation of morpholine by Mycobacterium aurum strain MO1 (Combourieu et al. 1998). 2. The degradation of phenylacetate has remained enigmatic for several years, and details of the pathway used by Escherichia coli were elucidated using 13C-labeled substrate. Using the full complement of NMR technology, the structures of critical intermediates were determined and provided details of an unusual pathway (Ismail et al. 2003). 3. The degradation of acenaphthene is initiated by benzylic monooxygenation, and the pathway was determined using [1-13C]acenaphthene by the isolation of intermediate metabolites (Selifonov et al. 1998). Importantly, the method proved applicable even when only limited biotransformation of the substrates had taken place by partial oxidation. 4. The complex pathway for the anaerobic degradation of propionate by Smithella propionica and Methanospirillum hungatei involves reaction of two molecules of propionate followed by rearrangement to 3-ketohexanoate. The details were elucidated using 13C-propionate labeled at C1, C2, C3, or at both C2 and C3 (de Bok et al. 2001). The procedure is particularly suited to the study of anaerobic transformations since there are no problems resulting from problems with oxygen limitation. A good example is provided by the application of 13C NMR to the intricate relations of fumarate, succinate, propionate, and acetate in a syntrophic organism both in the presence (Houwen et al. 1991), and in the absence of methanogens (Plugge et al. 1993; Stams et al. 1993). A single example will be given to illustrate both the strengths and the limitations of the technique. During the metabolism of 2-13C-acetate in methylotrophic strains of Pseudomonas sp., it was shown that the substrate was converted into α,α-trehalose in isocitrate lyase-negative strains, though not in one that synthesized this enzyme. In addition, an unknown compound was revealed by the in vitro experiments, but was not present in perchloric acid extracts of the cells. Possibly more disturbing, however, was the fact that the analysis with a strain during growth with 13C-methanol did not reveal the presence of intermediates known to be part of the serine pathway that functions in this organism (Narbad et al. 1989). Nitrogen 15N Apart from its use in studies of nitrogen fixation, 15N-labeled substrates have been used to assess the association of contaminants with organic matter in soil, and in metabolism.

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Association of Xenobiotics with Soil Organic Material N NMR has been used in conjunction with 13C NMR in studies on reactions of [15N] hydroxylamine with fulvic and humic substances (Thorn et al. 1992). 2. The availability of 15N aniline has made possible a direct study of the reactions of aniline with humic and fulvic acids (Thorn et al. 1996), and the detection of resonances attributed to anilinoquinone, imines, and N-heterocyclic compounds are fully consistent with reactions involving quinone and ketone groups. 3. The 15N-labeled amines produced by partial and total reduction of the nitro groups in 2,4,6-trinitrotoluene reacted with carbonyl groups (quinones and ketones) in humic acid to produce a range of products (Thorn and Kennedy 2002). 1.

15

Metabolism 1. The metabolism of methylamine by Pseudomonas sp. strain MA was studied using combined 13C and 15N NMR (Jones and Bellion 1991). This organism uses a complex pathway for the degradation of methylamine involving N-methylglutamate and γ-glutamylmethylamide, and these could be identified at different rates of oxygenation. 2. The degradation of benzothiazole was examined in two strains of Rhodococcus. The structure of the metabolite as 2,6-dihydroxybenzothiazole was determined by long-range 1 H–15N gradient heteronuclear multiple-bond correlation without 15N enrichment (Besse et al. 2001). 3. The metabolism of 2,4,6,8,10,12-hexanitro-2,4,6,8,10,12-hexaazaisowurtzitane (CL-20) labeled in the ring atoms with 15N was examined using salicylate 1-monooxygenase from Pseudomonas sp. strain FA1, and it was possible to propose a pathway (Bhushan et al. 2004). Oxygen 17O Application of 17O NMR has been made to a restricted range of chlorinated aromatic compounds (Kolehmainen et al. 1992), and has been used to establish the source of oxygen in the metabolites produced from 17O acetate and 17O2 by Aspergillus melleus (Staunton and Sutkowski 1991). Fluorine 19F Application to fluorine-containing molecules is particularly attractive since naturally occurring fluorine is monoisotopic, and the range of chemical shifts in fluorine compounds is very much greater than the proton shifts for hydrogen-containing compounds. Although only a few examples are given as illustration, there is a vast potential for the application of 19F NMR to metabolic studies of fluorine compounds. The principles and applications have been given in a substantive review by Stanley (2002). 1. Substantial attention has been devoted to the metabolism of 5-fluorouracil and related compounds. For example, 19F NMR was used successfully both in cell extracts and in whole mycelia to elucidate anabolic reactions involving pyrimidine nucleotides and degradation to α-fluoro-β-alanine in the fungus Nectria haematococca (Parisot et al. 1989, 1991). 2. Fluoroacetate and 4-fluorothreonine are synthesized from fluoride by Streptomyces cattleya, and analysis of supernatants was used to elucidate the details of their biosynthesis. They were apparently synthesized by independent routes, and it was suggested that what is formally glycolate could be their precursor (Reid et al. 1995). 3. Oxidative decarboxylation of hydroxybenzoates by the yeast Candida parapsilosis is catalyzed by a flavin monooxygenase that is able to use a range of fluorinated hydroxybenzoates that were examined by 19F NMR (Eppink et al. 1997).

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F NMR was used to study the effect of pH on the hydroxylation of 3-fluorophenol by the hydroxylase from Trichosporon cutaneum, and revealed that the ratio of products as well as the yield was pH-dependent (Peelen et al. 1993). 5. The metabolism of 3,5- and 2,5-difluorobenzoate was studied in a mutant of Pseudomonas putida PpJT103 that is unable to aromatize the 1,2-dihydrodiol of benzoate. 19F NMR was used to establish that in both 3,5- and 2,5-difluorobenzoate, dioxygenation took place at C1 and C2, and that for the latter the dihydrodiol lost fluoride to produce 4-fluorocatechol (Cass et al. 1987). 6. The metabolism of fluorophenols by phenol hydroxylase from Trichosporium cutaneum, catechol 1,2-dioxygenase from Pseudomonas arvilla strain C-1, and by the fungus Exophilia jeanselmei has been examined, and detailed NMR data were given for the ring fission fluoromuconates (Boersma et al. 1998). This methodology is worthy of wider exploitation in the study of other groups of organofluorine compounds that are of industrial importance as agrochemicals and pharmaceuticals and fluorinated compounds such as polyfluorinated octanoates and sulfonates that have awakened increasing environmental interest. Phosphorus 31P Until recently, there has been only limited interest in the catabolism of organophosphorus compounds, although very considerable attention has been directed to anabolic reactions. One example is given to illustrate the strengths and limitations of the technique. 31P was used to examine the effect of ethanol on the metabolism of glucose by Zymomonas mobilis. Whereas the sensitivity was sufficient to establish changes in nucleoside triphosphates during in vivo experiments, details of the changes in the various phosphorylated metabolites necessitated the use of perchloric acid cell extracts (Strohhäcker et al. 1993). Illustrative examples include the following: 1. Nucleotide pools and transmembrane potential in bacteria after exposure to pentachlorophenol were investigated using 31P NMR. Differences were used to differentiate Escherichia coli, which does not degrade this substrate and a Flavobacterium sp., which is able to do so (Steiert et al. 1988). 2. Two strains of Sphingomonas sp. that could degrade pentachlorophenol maintained their levels of ATP even in the presence of high concentrations of pentachlorophenol. Analysis of the lipids using 31P NMR showed that this could be attributed to the increased levels of cardiolipin (Lohmeier-Vogel et al. 2001). Silicon 29Si 29

Si NMR of hexamethyldisiloxane has been examined using a 750-MHz 1H resonance frequency, and a number of technical issues were discussed (Knight and Kinrade 1999). The use of chromiumIII acetylacetonate was used to reduce the 29Si relaxation time with a DEPT-45 pulse sequence. This indicated the potential application to the metabolism of organosilicon compounds.

APPLICATION OF EPR Compared with the extensive application of NMR procedures, electron paramagnetic resonance (EPR) is used less frequently except for analysis of the state of metals in enzymes and coenzymes. There has, however, been increasing awareness of the role of radical-mediated reactions and some examples are used to illustrate its potential.

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Elucidation of the Unexpected Complexity of an Apparently Straightforward Metabolic Pathway a. The metabolism of glycerol trinitrate by Phanerochaete chrysosporium does not proceed by hydrolysis to nitrate and glycerol, but involves formation of nitric oxide bound to both nonheme and heme proteins (Servent et al. 1991). b. The degradation of tetrachloromethane mediated by the bacterial metabolite pyridine 2,6-dithiocarboxylate in the presence of Cu(II) is initiated by one-electron reduction of the substrate by the copper complex. This was identified by EPR using 13CCl4 (Lewis et al. 2001). Hydroxyl Radicals and Their Role in the Toxicity of H2O2 to Bacteria In a study directed to the analysis of the role of Fe and the generation of H2O2 in Escherichia coli (McCormick et al. 1998), hydroxyl radicals were specifically trapped by reaction with ethanol to give the α-hydroxyethyl radical. This formed a stable adduct with α-(4-pyridyl-1-oxide)-N-t-butyl nitroxide that was not formed either by superoxide or hydroxyl radicals. The role of redox-reactive iron is to use EPR to analyze the EPR-detectable ascorbyl radicals. Elucidation of the Mechanism of Pyruvate Formate Lyase This enzyme catalyzes the conversion of pyruvate to formate and acetyl CoA and is a key enzyme in the anaerobic degradation of carbohydrates in some Enterobacteriaceae. Using an enzyme selectively 13C-labeled with glycine, it was shown by EPR that the reaction involves production of a free radical at C-2 of glycine (Wagner et al. 1992). This was confirmed by destruction of the radical with O2, and determination of part of the structure of the small protein that contained an oxalyl residue originating from gly-734. The Function of Humic Acids in Reactions Catalyzed by Geobacter metallireducens This organism is able to oxidize acetate to CO2 under anaerobic conditions in the presence of Fe(III). A study of the intermediate role of humic and fulvic acids used ESR to detect and quantify free radicals produced from oxidized humic acids by cells of G. metallireducens in the presence of acetate. There was a substantial increase in the radical concentration after incubation with the cells, and it was plausibly suggested that these were semiquinones produced from quinone entities in the humic and fulvic structures (Scott et al. 1998). The Mechanism of Anaerobic Activation of Toluene Benzylsuccinate is synthesized during the anaerobic degradation of toluene under denitrifying conditions. It involves activation of toluene to form a benzyl radical by a mechanism involving glycyl radical catalysis that has been confirmed and analyzed using EPR (Krieger et al. 2001; Leuthner et al. 1998; Verfürth et al. 2004). Dioxygenation with Elimination of Halide Burkholderia cepacia strain 2CBS is able to degrade ortho-halogenated benzoates by dioxygenation to catechol with the elimination of halide and decarboxylation. The enzyme contains a ferredoxinand a Rieske-type [2Fe–2S] center. These could be distinguished on the basis of their EPR spectra, and the results were compared with those for other [2Fe–2S] clusters (Riedel et al. 1995).

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The Anaerobic Degradation of Benzoate Benzoyl-CoA reductase carries out the two-electron reduction of the aromatic ring during the anaerobic degradation of benzoate by Thauera aromatica. This involves two-electron transfer from ferredoxin, and a combination of EPR and Mössbauer spectroscopy showed the presence of three different clusters, while inactivation by oxygen was associated with partial conversion of [4Fe–4S] clusters to [3Fe–4S] clusters (Boll et al. 2000). Manganese-Containing Enzymes a. Manganese-containing dioxygenase. Exceptionally, in Bacillus sp. strain JF8 that has been studied for its ability to degrade PCBs, the 2,3-dihydroxybiphenyl 1,2-dioxygenase contains manganese instead of iron. Its presence was confirmed by EPR for which signals at g = 2.02 and g = 4.06 having the sixfold splitting characteristic of Mn(II) were observed (Hatta et al. 2003). b. The acetone carboxylase in Rhodobacter capsulatus catalyzes the conversion of acetone to acetoacetate at the expense of ATP. Acetone-grown cells had an EPR signal centered at g = 2, which could be associated with the enzyme since this was absent in malate-grown cells (Boyd et al. 2004). Turnover of Naphthalene Dioxygenase Details of the mechanism of naphthalene dioxygenase during a single turnover of the enzyme have been revealed, and confirmed the separate roles of the dioxygenase and the ferredoxin electron transfer protein. This made it possible to propose a reaction cycle for the reaction (Wolfe et al. 2001).

ENZYME STRUCTURE Several factors have contributed to elucidating enzymological details of degradative pathways: 1. Advances in molecular biology have led to the construction of strains that include the relevant degradative genes. Amplification of these, and expression of them in Escherichia coli has made possible the isolation of pure enzymes in quantities suitable for crystallization. 2. There has been remarkable progress in techniques for crystallizing proteins. 3. Advances in the technology of x-ray diffraction have made it possible to achieve threedimensional structures of enzymes together bound, in many cases, to their substrates. Illustrative examples include the following: 1. Degradation of halogenated substrates: a. The structure of the active site of haloalkane dehalogenase from Xanthobacter autotrophicus strain GJ10 with 1,2-dichloroethane as substrate was examined at different stages of the reaction that provided intricate details of the mechanism of the reaction (Verschuren et al. 1993). b. The structure of the reaction intermediate of L-2-chloroacid dehalogenation from Pseudomonas sp. strain YL showed that the enzyme functions through the formation of an ester intermediate with Asp-10 (Li et al. 1998). 2. Metalloenzymes: a. The structure of the unusual copper-containing dioxygenase in Aspergillus japonicus bound the natural substrate quercitin and kaemferol to it and made it possible to provide a mechanism for its activity (Steiner et al. 2002).

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b. The structure of the carbon monoxide dehydrogenase from the anaerobe Carboxydothermus hydrogenoformans revealed an unusual cluster of metal atoms in an asymmetric [Ni–4Fe–4S] cluster (Dobbek et al. 2001). c. Methyl-coenzyme M reductase plays a key role in methanogenesis, and the structure of the enzyme from Methanobacterium thermoautrophicum made it possible to assign a specific role to coenzyme B and the nickel porphinoid coenzyme F430 in methanogenesis (Ermler et al. 1997). d. The structure of the copper quinoprotein amine oxidase, which is used by a number of Enterobacteriaceae for deamination of primary amines, revealed significant features including the role of Cu(II) that is reduced to Cu(I), and the catalytic role of Asp383 (Wilmot et al. 1999).

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Parisot D, MC Malet-Martino, R Martino, P Crasnier (1991) 19F nuclear magnetic resonance analysis of 5-fluorouracil metabolism in four differently pigmented strains of Nectria haematococca. Appl Environ Microbiol 57: 3605–3612. Peelen S, IMCM Rietjens, WJH van Berkel, WAT van Workum, J Vervoort (1993) 19F-NMR study on the pHdependent regioselectivity and rate of the ortho-hydroxylation of 3-fluorophenol by phenol hydroxylase from Trichosporon cutaneum. Eur J Biochem 218: 345–353. Plugge CM, C Dijkema, AJM Stams (1993) Acetyl-CoA cleavage pathway in a syntrophic propionate oxidizing bacterium growing on fumarate in the absence of methanogens. FEMS Microbiol Lett 110: 71–76. Poupin P, N Truffaut, B Combourieu, M Sancelelme, H Veschambre, AM Delort (1998) Degradation of morpholine by an environmental Mycobacterium strain involves a cytochrome P-450. Appl Environ Microbiol 64: 15–165. Pusecker K, J Schewitz, P Gfrörer, L-H Tseng, K Albert, E Bayer (1998) On-line coupling of capillary electrochromatography, capillary electrophoresis, and capillary HPLC with nuclear magnetic resonance spectroscopy. Anal Chem 70: 3280–3285. Reid KA, JTG Hamilton, RD Bowden, D O’Hagan, L Dasaradhi, MR Amin, DB Harper (1995) Biosynthesis of fluorinated secondary metabolites by Streptomyces cattleya. Microbiology (UK) 141: 1385–1393. Riedel A, S Fetzner, M Rampp, F Lingens, U Liebl, J-L Zimmermann, W Nitschke (1995) EPR, electron spin echo envelope modulation, and electron nuclear double resonance studies of the 2Fe–2S centers of the 2-halobenzoate 1,2-dioxygenase from Burkholderia (Pseudomonas) cepacia 2CBS. J Biol Chem 270: 30869–30873. Scott DT, DM McKnight, EL Blunt-Harris, SE Kolesar, DR Lovley (1998) Quinone moieties act as electron acceptors in the reduction of humic substances by humics-reducing microorganisms. Environ Sci Technol 32: 2984–2989. Selifonov SA, PJ Chapman, SB Akkerman, JE Gurst, JM Bortiatynski, MA Nanny, PG Hatcher (1998) Use of 13C nuclear magnetic resonance to assess fossil fuel biodegradation: fate of [1-13C]acenaphthene in creosote polycyclic aromatic compound mixtures degraded by bacteria. Appl Environ Microbiol 64: 1447–1453. Servent D, C Ducrorq, Y Henry, A Guissani, M Lenfant (1991) Nitroglycerin metabolism by Phanerochaete chrysosporium: evidence for nitric oxide and nitrite formation. Biochim biophys Acta 1074: 320–325. Sidelmann UG, U Braumann, M Hofmann, M Spraul, JC Lindon, JK Nicholson, SH Hansen (1997) Directly coupled 800 MHz HPLC-NMR spectroscopy of urine and its application to the identification of the major phase II metabolites of tolfenamic acid. Anal Chem 69: 607–612. Small FJ, SA Ensign (1995) Carbon dioxide fixation in the metabolism of propylene and propylene oxide by Xanthobacter strain Py2. J Bacteriol 177: 6170–6175. Stams AJM, JB van Dijk, C Dijkema, CM Plugge (1993) Growth of syntrophic propionate-oxidizing bacteria with fumarate in the absence of methanogenic bacteria. Appl Environ Microbiol 59: 1114–1119. Stanley PD (2002) Principles and topical applications of 19F NMR spectrometry. Handbook Environ Chem 3N: 1–56. Staunton J, AC Sutkowski (1991) 17O NMR in biosynthetic studies: aspyrone, asperolactone isoasperoloactone, metabolites of Aspergillus melleus. J Chem Soc Chem Commun 1106–1108. Steiert JG, WJ Thoma, K Ugurbil, RL Crawford (1988) 31P nuclear magnetic resonance studies of effects of some chlorophenols on Escherichia coli and a pentachlorophenol-degrading bacterium. J Bacteriol 170: 4954–4957. Steiner RA, KH Kalk, BW Dijkstra (2002) Anaerobic enzyme substrate structures provide insight into the reaction mechanism of the copper-dependent quercitin 2,3-dioxygenase. Proc Natl Acad USA 99: 16625–16630. Strohhäcker J, AA de Graaf, SM Schoberth, RM Wittig, H Sahm (1993) 31P nuclear magnetic resonance studies of ethanol inhibition in Zymomonas mobilis. Arch Microbiol 159: 484–490. Subramanian R, AG Webb (1998) Design of solenoidal microcoils for high-resolution 13C NMR spectroscopy. Anal Chem 70: 2454–2458 . Thorn KA, JB Arterburn, MA Mikita (1992) 15N and 13C NMR investigation of hydroxylamine-derivatized humic substances. Environ Sci Technol 26: 107–116. Thorn KA, KR Kennedy (2002) 15N NMR investigation of the covalent binding of reduced TNT amines to soil humic acid, model compounds, and lignocellulose. Environ Sci Technol 36: 3787–3796. Thorn KA, PJ Pettigrew, WS Goldenberg, EJ Weber (1996) Covalent binding of aniline to humic substances. 2. 15N NMR studies of nucleophilic addition reactions. Environ Sci Technol 30: 1764–1775.

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Unkefer CJ, RE London (1984) In vivo studies of pyridine nucleotide metabolism in Escherichia coli and Saccharomyces cerevisiae by carbon-13 NMR spectroscospy. J Biol Chem 2311–2320. Verfürth K, AJ Pierik, C Leutwein, S Zorn, J Heider (2004) Substrate specificities and electron paramagnetic resonance properties of benzylsuccinate synthesis in anaerobic toluene and m-xylene metabolism. Arch Microbiol 181: 155–162. Verschuren KHG, F Seljée, HJ Rozeboom, KH Kalk, BW Dijkstra (1993) Crystallographic analysis of the catalytic mechanism of haloalkane dehalogenase. Nature 363: 693–698. Wagner AFV, M Frey, FA Neugebauer, W Schäfer (1992) The free radical in pyruvate formate-lyase is located on glycine-734. Proc Natl Acad USA 89: 996–1000. Wilmot CM, J Hajdu, MJ Mc Pherson, PF Knowles, SEV Phillips (1999) Visualization of dioxygen bound to copper during enzyme catalysis. Science 286: 1724–1728. Wolfe MD, JV Parales, DT Gibson, JD Lipscomb (2001) Single turnover chemistry and regulation of O2 activation by the oxygenase component of naphthalene 1,2-dioxygenase. J Biol Chem 276: 1945–1953.

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Section III Pathways and Mechanisms of Degradation and Transformation

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Aliphatic Compounds

PART 1 ALKANES AND RELATED COMPOUNDS Petroleum hydrocarbons are used as automotive fuels and as monomers for the production of a range of plastics. They provide the basis of the petrochemical industry and the halogenated derivatives, which are discussed in Part 2 of this chapter.

ALKANES There is an enormous literature on the microbial degradation of alkanes; this has been motivated by applications as diverse as the utilization of methane and methanol for the production of singlecell protein or to combating oil spills. The number and range of microorganisms is impressive, and includes many different taxa of bacteria, yeasts, and fungi, including thermophilic species of Geobacillus (Nazina et al. 2001). Extensive reviews that cover most aspects have been presented (Ratledge 1978; Britton 1984). Aerobic Conditions The simplest alkane is methane, but the pathways for its degradation and assimilation do not reflect this structural simplicity. In outline, the pathway of degradation is straightforward and involves three stages: 1. Oxidation to methanol catalyzed by methane monooxygenase (MMO). 2. Conversion of methanol into formaldehyde by methanol dehydrogenase. A complex array of genes is involved in this oxidation and the dehydrogenase contains pyrroloquinoline quinone (PQQ) as a cofactor (references in Ramamoorthi and Lidstrom 1995). Details of its function must, however, differ from that of methylamine dehydrogenase that also contains a quinoprotein—tryptophan tryptophylquinone (TTQ). 3. Further degradation to formate may involve the tetrahydrofolate (THF) or tetrahydromethanopterin pathways (Marx et al. 2003) that are comparable to those used for methyl chloride and methyl bromide, which are discussed in Part 3 of this chapter. In addition, the cells must be capable of synthesizing cell material from the substrate, so that some fraction of the C1 metabolites must also be assimilated. Several distinct pathways for this have been described, but these are merely summarized here since a comprehensive and elegant presentation of the details has been given (Anthony 1982): a. The ribulose bisphosphate pathway for the assimilation of CO2 that is identical to the Benson–Calvin cycle used by photosynthetic organisms. b. The ribulose monophosphate cycle for the incorporation of formaldehyde. c. The serine pathway for the assimilation of formaldehyde. A number of bacterial strains with monooxygenase activity have been described, and different types of MMO have played important roles in the degradation of a range of aliphatic compounds. MMO may exist in either a soluble (sMMO) form that has been more extensively studied or in a particulate (pMMO) form. These forms display different substrate ranges and different rates of transformation rates, and most methanotrophs express only the particulate form of the enzyme (Hanson and Hanson 1996).

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The soluble MMO from both type I and type II methanotrophs is a three-component system. It comprises (i) a nonheme iron hydroxylase containing an oxo-bridged binuclear Fe cluster, (ii) a metal-free protein component without redox cofactors, and (iii) an NADH reductase containing FAD and a [2Fe–2S] cluster (Fox et al. 1989). The particulate enzyme contains copper or both copper and iron, and the concentration of copper determines the catalytic activity of the enzyme (Sontoh and Semrau 1998). The enzyme consists of three components, a hydroxylase, a regulatory protein that is not directly involved in electron transfer between the hydroxylase, and a third protein that is a reductase containing FAD and a [2Fe–2S] cluster. Details of the structure of the hydroxylase and the mechanism of its action involving the FeIII–O–FeIII at the active site have been given in a review (Lipscomb 1994). An additional aspect is the wide spectrum of substrates that can be metabolized by the MMO system. These include methyl fluoride and further examples, which are given in Figure 7.1. Attention is drawn to the similarity of this enzyme to the monooxygenase involved in the oxidation of ammonia, while the broad substrate specificity of cyclohexane oxygenase has been noted in Chapter 2, Figure 2.10. Microorganisms are able to degrade a structurally wide range of higher alkanes. These include linear alkanes with both even numbers of carbon atoms up to at least C30, some odd-numbered alkanes including the plant wax C29H60 (Hankin and Kolattukudy 1968), and branched alkanes such as pristane (2,6,10,12-tetramethylpentadecane) (McKenna and Kallio 1971; Pirnik et al. 1974). Degradation is initiated by hydroxylation and several mechanisms for carrying this out have been demonstrated. In addition, strains differ in their specificity for the chain length of the hydrocarbon. Pseudomonas putida (P. oleovorans) Gpo1 carries the OCT plasmid and hydroxylates alkanes of chain length C3–C4 and C6 –C14, and ω-oxidation of fatty acids from C6 to C12 (McKenna and Coon 1970). In contrast, the hydroxylase from Acinetobacter sp. strain M-1 is able to accept alkanes up to C30 and maximally of C12 (Maeng et al. 1996). A mechanistic peculiarity of this strain is noted later.

CH3(CH2)6−CH2OH

CH3(CH2)6−CH3

CH3 Br

CH3(CH2)5−CH(OH)CH3

CH 2 O

;

O CCH3−CH

;

CH3CH2−O−CH2CH3

CH2

CH3−CH

CH2

CH3CH2OH

OH ; N

N O

CH3

CH3

CH2OH

CO2H

OH

FIGURE 7.1

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Examples of the reactions catalyzed by methane monooxygenase.

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299

The ability to use short-chain alkanes (C2–C4) is found among both Gram-positive organisms belonging to the Corynebacterium–Nocardia–Mycobacterium–Rhodococcus group, for example, Mycobacterium vaccae strain JOB5 (Hamamura et al. 1999), and (C3 –C4) in the Gram-negative Pseudomonas putida Gpo1 (Johnson and Hyman 2006). In Nocardioides sp. strain CF8, there appear to be two different hydroxylases—one inducible by butane for C2–C6 alkanes, and another for >C6 alkanes (Hamamura et al. 2001). Two reaction sequences are involved in the bacterial degradation higher n-alkanes: 1. Terminal hydroxylation followed by successive dehydrogenations to the corresponding alkanol, aldehyde, and alkanoic acid. These are then degraded by β-oxidation to yield ultimately acetate from even-membered alkanes or propionate from odd-membered alkanes (Figure 7.2). Degradation of propionate may occur by several pathways (Wegener et al. 1968): (a) the methylmalonate, (b) the acrylate, and (c) the 2-methylcitrate (Figure 7.3). The last of these is initiated by reaction of propionate with oxalacetate, followed by formation of 2-methylisocitate from which succinate is produced with loss of pyruvate. This has been established in enteric organisms (Textor et al. 1997; Horswill and EscalanteSemerena 1999), in Corynebacterium glutamicum (Claes et al. 2002), and in Ralstonia eutropha (Brämer and Steinbüchel 2001). 2. Subterminal hydroxylation followed by oxidation to ketones. These are then degraded by a sequence involving Baeyer–Villiger monooxygenation to lactones, hydrolysis with loss of a C2 unit, and the reactions following terminal hydroxylation (Figure 7.4). An additional pathway is given in the section on the degradation of alkanones.

+

FIGURE 7.2

Terminal oxidation of alkanes. CH3−CH2−CO2H (a)

CO 2 H

+

(c)

(b)

CH2

HO2CCH2COCO2H

CH3−CH−CO2H

CH−CO 2H

CH3−CH−CO2H

HO−C−CO2H HO−CH2−CH 2CO2H

CH2CO2H

CH3−CH(OH)CO2H

CH2CO2H CH2CO2H

CH2CO2H CH2CO2H

+ CH3COCO2H

FIGURE 7.3 Alternative pathways for the aerobic degradation of propionate. RCH2CH2CH3 RCH2O−COCH3

FIGURE 7.4

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RCH2COCH3

RCH2CH(OH)CH3 RCH2OH

RCO2H

Subterminal oxidation of alkanes.

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There are important additional features: a. For long-chain alkanes, hydroxylation may occur at both ends of the chain with the production of α,ω-dicarboxylic acids that are then degraded by β-oxidation. For example, the oxidation of pristane proceeds by both β-oxidation and ω-oxidation (McKenna and Kallio 1971; Pirnik et al. 1974) (Figure 7.5). The production of α,ω-dicarboxylic acids can depend critically on the chain length of the hydrocarbon. Although cells of Gordonia (Corynebacterium) sp. strain 7E1C produce α,ω-dodecanedioic acid during growth with dodecane, the analogous α,ω-hexadecanedioic acid is not produced from hexadecane (Broadway et al. 1993). The ω-hydroxylase in Pseudomonas putida (oleovorans) consists of three components, a soluble NADH-rubredoxin reductase, a soluble rubredoxin, and an integral-membrane hydroxylase, analogous to the enzymes involved in α-hydroxylation (Shanklin et al. 1997). It has been reported that the enzyme can also carry out epoxidation (May and Abbott 1973), and that the ratio of the hydroxylase and epoxidase remained essentially constant during purification (Ruettinger et al. 1977). b. The existence of chain branching may present an obstacle to degradation by preventing β-oxidation. This can be circumvented in a carboxylation pathway. The critical step is carboxylation of the methyl group at C3 (Fall and Hector 1977) followed by hydration of the unsaturated coenzyme A ester, oxidation, and loss of –CH2CO2H that is followed by hydrolytic fission of the β-ketoester (Figure 7.6). These reactions are analogous to those involved in the degradation of the acyclic monoterpenoids citronellol and geraniol by strains of Pseudomonas aeruginosa. After dehydrogenation of the alkanols, degradation of the carboxylates involves the coenzyme A esters (Díaz-Pérez et al. 2004; Höschle et al. 2005). The product is then fed into the l-leucine degradative pathway that is noted later. Although the pathways for degradation of citronellol and garaniol in Pseudomonas

CO 2 H

CO 2 H

HO 2 C

CO 2 H

HO 2 C

CO 2 H

CO 2 H CH 3 CO 2 H

FIGURE 7.5 Pathways for the biodegradation of pristane.

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Aliphatic Compounds

301

CH2OH CO2H

CO2H

FIGURE 7.6

O

CO2H

CO2H

Carboxylation pathway for the biodegradation of a branched alkane.

CH 3 HO2 C CH 3

CH 3

CH 3

CO2 H

CH 3

CH 3

CO2 H

CH 3

CH

CO2 H

CH 3 CH3 CH 2

FIGURE 7.7

CO2H

Degradation of 2,2-dimethylpropionate.

aeruginosa merge with the synthesis of isohexenyl glutaconyl-CoA, there is an important difference. Molybdenum is required in the oxidation of geraniol to geranate though not for the corresponding oxidation of citronellol to citronellate (Höschle and Jendrossek 2005). The carboxylation pathway is also used by Marinobacter sp. strain CAB for the aerobic degradation of 6,10,14-trimethylpentadecan-2-one (Rontani et al. 1997). c. Metabolites with quaternary carbon may be produced by degradation (Catelani et al. 1977). Although such compounds are generally recalcitrant, their degradation has been described, for example, dimethylpropionate produced from the aerobic degradation of tert-butylbenzene and 2,2-dimethylheptane. Dimethylmalonate (Kniemeyer et al. 1999) and pivalate (Probian et al. 2003) can be mineralized under denitrifying conditions by reactions that may involve corrin-mediated rearrangements (Ratnatilleke et al. 1999) (Figure 7.7). The analogous rearrangement of bisphenol-A produces 4-hydroxyacetophenone and 4-hydroxybenzoate (Lobos et al. 1992; Spivack et al. 1994) after fission of the hydroxylated intermediates that are produced in Sphingomonas sp. strain AO1 by a cytochrome P450 (Sasaki et al. 2005) (Figure 7.8). d. Although the degradation of alkylphenols can be initiated by oxidation of the alkyl groups to carboxylic acids, the degradation of nonylphenol isomers with quaternary side chains in strains of Sphingomonas sp. displays unusual features (Corvini et al. 2004; Gabriel et al. 2005). This may be described formally as oxygenation at the position para to the phenol group, followed by either of two rearrangements: (a) migration of the alkyl group to the C-3 position of the phenolic ring with formation of a hydroquinone (Figure 7.9a) or (b) insertion of the oxygen from the hydroxyl group between the alkyl group and the phenolic ring followed by oxygenation with loss of the tertiary alkanol and formation of benzoquinone (Figure 7.9b). Against this background it is appropriate to add a few details of the enzymology. Alkane hydroxylase activity is distributed among a number of Gram-negative bacteria and sequence analysis of proteins shows that these differ widely even among pseudomonads (Smits et al. 2002). The alkane hydroxylase system for Pseudomonas putida (P. oleovorans) GPo1, which has been most extensively studied, is composed of three components—an integral-membrane alkane hydroxylase (AlkB), a soluble

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Environmental Degradation and Transformation of Organic Chemicals CH 3

CH 3

CH 3

HO

OH

CH 2 OH

HO

OH

OH

OH CH 2 OH

HO

CH 3

HO

OH

CO 2 H

COCH 2 OH

OH

OH

CH 3

HO

FIGURE 7.8

CH 2 OH

HO

CO 2 H

COCH 3

OH

OH

Degradation of bisphenol-A.

OH

OH

H C

C OH

O

O

OH

(a)

O C

OH

C

O

(b)

OH

O

O C

HO

FIGURE 7.9

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OH O

O

O

C

OH + O

C

Biotransformations of nonylphenol.

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Aliphatic Compounds

303

rubredoxin (AlkG), and a soluble NADH-rubredoxin reductase (AlkT). The first of these is a representative of a large nonheme diiron cluster (Shanklin et al. 1997), while the others are involved in electron transport. The alkane hydroxylase of Acinetobacter sp. strain ADP1 is formally similar in requiring three components (Ratajczak et al. 1998). The alkane hydroxylase belongs to a family of nonheme iron oxygenases. There is some structural similarity between the nucleotide sequence of the integral-membrane alkane hydroxylase and the subunits of the monooxygenase encoded by xylA and xylM in the TOL plasmid that are involved in hydroxylation of the methyl groups in toluene and xylene in Pseudomonas putida PaW1 (Suzuki et al. 1991). In Pseudomonas putida ppG1 that carries the OCT plasmid, there is duplication of some of the loci. Those for alkane hydroxylation (alkA, alkB, and alkC) and for alkanol dehydrogenation (alcO) occur on the plasmid, whereas those for alcA and alcB, and for aldehyde dehydrogenation (aldA, aldB) occur in the chromosome (Grund et al. 1975). Alternative mechanisms for alkane degradation have been proposed for several organisms: 1. Alkane oxidation via a hydroperoxide was suggested many years ago, and seems to be operative in Acinetobacter sp. strain M-1 that has, in addition, a rather unusual range of substrates that include both n-alkanes and n-alkenes. The purified enzyme contains FAD and requires copper for activity (Maeng et al. 1996). 2. The operation of cytochrome P450 in alkane oxidation has been reported both in bacteria and in yeasts. It has been shown that alkane hydroxylases of CHYP 153 are widespread both in Gram-negative and Gram-positive bacteria that lack the integral membrane alkane hydroxylase (van Beilin et al. 2006). a. Cell extracts of Gordonia (Corynebacterium) sp. strain 7E1C oxidized n-octane to n-octanol, and the activity was resolved into a flavoprotein and a cytochrome P450 (Cardini and Jurtshuk 1968). Whereas cytochromes b and o were present in Acinetobacter calcoaceticus strain 69-V during growth with n-hexadecane, cytochrome P450 was found in other strains during growth with the same substrate (Asperger et al. 1981). b. Hydroxylation at the subterminal (ω-1, ω-2, and ω-3) positions of carboxylic acids with chain lengths of 12–18-carbon atoms—less readily of the corresponding alcohols and not at all with the alkanes—has been observed with a soluble enzyme system from a strain of Bacillus megaterium (Miura and Fulco 1975). Whereas in this organism, ω-2-hydroxylation is carried out by a soluble cytochrome P450BM-3 that is inducible with barbiturates (Narhi and Fulco 1987), ω-hydroxylation in Pseudomonas oleovorans that carries the OCT plasmid is mediated by a three-component hydroxylase that behaves like a cytoplasmic membrane protein (Ruettinger et al. 1974; Kók et al. 1989a,b). c. The cytochrome P450 CYP153A6 from Mycobacterium sp. strain HXN-1500 carries out hydroxylation of alkanes from C6 to C11 including 2-methyloctane at the terminal positions (Funhoff et al. 2006). d. Yeasts of the genera Candida and Endomycopsis are able to degrade alkanes with chain lengths >4 (Käppeli 1986; Tanaka and Ueda 1993). The alkane hydroxylase of Candida tropicalis is located within the microsomes that contain cytochrome P450 and NADPH-cytochrome c reductase (references in Käppeli 1986). In contrast, degradation of the alkanoate is carried out in peroxisomes that contain the β-oxidation enzymes—alkanoate oxidase, enoyl-CoA hydratase, and 3-hydroxyacylCoA dehydrogenase. Yeasts of the genus Candida excrete α,ω-dicarboxylates during growth on n-alkanes as the carbon source, and the cytochrome P450 gene family (CYP52) in Candida tropicalis is a determinant of this transformation (Craft et al. 2003). 3. In the hydroxylation of n-propane to n-propanol by Nocardia paraffinicum (Rhodococcus rhodochrous) ATCC 21198, the ratio of hydrocarbon to oxygen consumed was 2:1 and this suggests that the reaction of two molecules of propane and one molecule of dioxygen

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Environmental Degradation and Transformation of Organic Chemicals R

FIGURE 7.10

CH2

CH3

R CH

CH2

R CH2 CH2OH

R CH2 CO2H

Biodegradation of an alkane under anaerobic conditions.

produced two molecules of alkanol (Babu and Brown 1984). This reaction is formally analogous to the oxidation of 2-nitropropane to acetone by flavoenzymes in Hansenula mrakii (Kido et al. 1978) and Neurospora crassa (Gorlatova et al. 1998). 4. An unusual pathway has been proposed for the degradation of n-decane to the carboxylic acid by Pseudomonas sp. strain 196Aa under anaerobic conditions, apparently in the absence of an electron acceptor. This involved initial dehydrogenation, followed by hydroxylation and dehydrogenation (Figure 7.10), and was supported by purification of an NAD-linked alkane dehydrogenase, an NADPH-linked alkene hydroxylase, and an FAD Fe-dependent alcohol dehydrogenase (Parekh et al. 1977). Possibly this organism is not, in fact, an aerobic pseudomonad, although the survival of Pseudomonas aeruginosa under anaerobic conditions by pyruvate fermentation has been observed (Schreiber et al. 2006). 5. It is worth noting that the aerobic degradation of hexadecane may be accomplished at quite low oxygen concentrations (Michaelsen et al. 1992). Anaerobic Conditions Considerable effort has been directed to the anaerobic degradation of alkanes. There has been renewed interest in the anaerobic degradation of methane since this represents an important sink for methane in anaerobic sediments where it is produced by methanogenesis. Although the organisms responsible have not hitherto been isolated in pure culture, the anaerobic oxidation of methane probably involves consortia of archaea and sulfate-reducing bacteria (Teske et al. 2002; Schouten et al. 2003; Girguis et al. 2005), and may possibly occur by reverse methanogenesis (Hallam et al. 2004). Genes for both anaerobic methane oxidation and methanogenesis were found together with those for aerobic methane oxidation at the surface of a single site in Japan (Inagaki et al. 2004). There is evidence for the anaerobic degradation of alkanes to CO2, plausibly under conditions of sulfate reduction. In experiments with sediment slurries from contaminated marine areas, 14CO2 was recovered from 14C-hexadecane (Coates et al. 1997), and was inhibited by molybdate that is consistent with the involvement of sulfate reduction. Under sulfate-reducing conditions 14CO2 was produced from 14C[14,15]octacosane (C28H58) (Caldwell et al. 1998). Different mechanisms have been elucidated for the anaerobic degradation of higher alkanes, and both occurred simultaneously in a sulfate-reducing consortium (Callaghan et al. 2006): 1. Reaction at the subterminal position with fumarate can be initiated by a free-radical reaction (Rabus et al. 2001). In a denitrifying organism that degraded n-hexane (Figure 7.11) (Wilkes et al. 2002) and a sulfate-reducing organism that degraded hexadecane (Callaghan et al. 2006), this is followed by a rearrangement analogous to that of succinate → malonate and sequential loss of acetate. Comparable reactions have been described for a sulfate-reducing enrichment culture (Davidova et al. 2005), and the sulfate-reducing Desulfatibacillum aliphaticivorans that can oxidize C13 to C18 alkanes (Cravo-Laureau et al. 2005). An analogous pathway is used for the degradation of ethylcyclopentane by a sulfate-reducing enrichment culture, which is discussed in Part 2 of this chapter. 2. An alternative involved carboxylation at C3 followed by loss of the C1–C2 fragment and further degradation of the resulting terminal carboxylate by β-oxidation (So et al. 2003) (Figure 7.12).

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Aliphatic Compounds

CH 3 (CH 2 )3 CH 2 CH 3

305 CH 3 CO 2 H

CH 3 (CH 2 )3 CH

CH 3

CO 2 H

CH 3 (CH 2 )3 CH−CH2CH COSCoA

CO 2 H CH 3 CH 3 (CH 2 )3 CH−CH

FIGURE 7.11

CH−COSCoA

CH 3 CO−SCoA

Anaerobic degradation of alkanes. CO 2 H

CH 3 (CH 2 )n CH 2 CH 2 CH 3

FIGURE 7.12

CH 3 (CH 2 )n CH

CH 2 CH 3

CH 3 (CH 2 )n CH 2 CO 2 H + C 2

Anaerobic degradation of n-alkanes.

ALKENES Aerobic Conditions Two kinds of investigations have been carried out (a) growth of microorganisms at the expense of alkenes and (b) biotransformations resulting in the synthesis of epoxides. For example, growth has been demonstrated at the expense of propene and butene (Van Ginkel and de Bont 1986). An interesting observation is the pathway for the degradation of intermediate n-alkenes produced by an aerobic organism under anaerobic conditions (Parekh et al. 1977). Although the generality of this pathway remains unknown, it is clearly possible that even aerobic bacteria under anoxic conditions might accomplish comparable degradations. Attention is drawn to the possibility that intermediate metabolites may be incorporated into biosynthetic pathways. For example, hexadecene is oxidized by the fungus Mortierella alpina by ω-oxidation (Shimizu et al. 1991), but the lipids contain carboxylic acids with both 18 and 20 carbon atoms including the unusual polyunsaturated acid 5-cis, 8-cis,11-cis,14-cis,19-eicosapentaenoic acid. Direct fission of the double bond by oxidation exemplifies the simplest, though less common reaction. A few illustrations are given: 1. The degradation of squalene by Marinobacter sp. strain 2sq31 (Rontani et al. 2002) is initiated by oxidative fission, although the subsequent steps are carried out by β-oxidation and carboxylation that are comparable to those used for branched alkanes. 2. An extracellular enzyme from Xanthomonas sp. is able to degrade poly(cis-1,4-isoprene) with the production of 12-keto-4,8-dimethyltrideca-4,8-diene-1-al (Braaz et al. 2004), and functions as a heme-dependent oxygenase (Braaz et al. 2005). The first step in the aerobic degradation of alkenes is, however, generally epoxidation. Alkene monooxygenase is closely related to the aromatic monooxygenases, and is able to hydroxylate benzene, toluene, and phenol (Zhou et al. 1999), while the alkane hydroxylase from Pseudomonas oleovorans is able to carry out both hydroxylation and epoxidation (Ruettinger et al. 1977). In Xanthobacter sp. strain Py2, which can carry out epoxidation of alkenes from C2 to C6, the monooxygenase has been characterized, and consists of an oxygenase, a ferredoxin, and a reductase (Small and Ensign 1997). Epoxidation may be followed by several alternatives, which may be summarized before further discussion: (a) hydrolysis to a 1,2-diol, (b) nucleophilic attack by coenzyme M followed by dehydrogenation, and (c) a reductive glutathione-mediated reaction. In detail, the degradation of epoxides is quite varied and several different pathways have been observed. 1. Degradation of epichlorohydrin (1-chloro-2,3-epoxypropane) may proceed by hydrolysis of the epoxide to 3-chloro-1,2-propanediol that is then converted successively into

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3-hydroxy-1,2-epoxypropane (glycidol) followed by hydrolysis to glycerol before degradation (van den Wijngaard et al. 1989). Epoxide hydrolases have been isolated and characterized from bacteria that are able to use epoxides as growth substrates. A Corynebacterium sp. is able to grow with alicyclic epoxides, and the sequence of the hydrolase (Misawa et al. 1998) is similar to the enzyme from Agrobacterium radiobacter strain AD1 that used epichlorohydrin (1-chloro-2,3-epoxypropane) as growth substrate (Rink et al. 1997). Examination of mutants of this strain prepared by site-directed mutagenesis showed that the mechanism involves nucleophilic attack by Asp107 at the terminal position of the substrate followed by hydrolysis of the resulting ester mediated by His275. Analogy may be noted with the inversion accompanying hydrolysis of 2-haloacids mediated by l-2-haloacid hydrolase. Limonene-1,2-epoxide hydrolases from Rhodococcus erythropolis DCL14 that are intermediates in the degradation of both (+)-(4R)- and (−)-(4S)-limonene (Van der Werf et al. 1999) differ from these groups of enzymes, and do not involve the catalytic function of histidine residues (Van der Werf et al. 1998). 2. Hydrolysis to the diol followed by dehydration to the aldehyde and oxidation to the carboxylic acid is used by a propene-utilizing species of Nocardia (de Bont et al. 1982). Although an ethene-utilizing strain of Mycobacterium sp. strain E44 degrades ethane-1,2-diol by this route, the diol is not an intermediate in the metabolism of the epoxide (Wiegant and de Bont 1980). 3. The aldehyde may also be produced directly from the epoxide. This occurs in the metabolism of ethene by Mycobacterium sp. strain E44 (Wiegant and de Bont 1980), of styrene by a strain of Xanthobacter sp. strain 124X (Hartmans et al. 1989), and by Corynebacterium sp. strains ST-5 and AC-5 (Itoh et al. 1997). The reductase in the coryneforms has a low substrate specificity, and is able to reduce acetophenone to 3-phenylethan-2-ol with an enantiomeric excess >96%. In Rhodococcus rhodochrous, however, styrene is degraded by ring dioxygenation with the vinyl group intact (Warhurst et al. 1994): 2-vinyl-cis, cis-muconate is produced by catechol 1,2-dioxygenase as a terminal metabolite and complete degradation is carried out by catechol 2,3-dioxygenase activity that is also present. 4. Xanthobacter sp. strain Py2 may be grown with propene or propene oxide. On the basis of amino acid sequences, the monooxygenase that produces the epoxide was related to those that catalyzes the monooxygenation of benzene and toluene (Zhou et al. 1999). The metabolism of the epoxide is initiated by nucleophilic reaction with coenzyme M followed by dehydrogenation (Figure 7.13a). There are alternative reactions, both of which are dependent on a pyridine nucleotide-disulfide oxidoreductase (Swaving et al. 1996; Nocek et al. 2002): a. In the absence of CO2, by transformation to acetone that is not further degraded. b. In the presence of CO2 by carboxylation to 3-ketobutyrate (acetoacetate) (Allen et al. 1999; Krum and Ensign 2001). Acetoacetate is used partly for cell growth and partly converted into the storage product poly-β-hydroxybutyrate (Small and Ensign 1995). Kinetic and 13C NMR experiments confirmed that acetoacetate was the first product from which β-hydroxybutyrate is formed as a secondary metabolite with acetone as the terminal metabolite (Allen and Ensign 1996). The epoxide carboxylase consists of a three-component enzyme—all three of which are necessary for activity (Allen and Ensign 1997). Coenzyme M (2-mercaptoethanesulfonate) is required, component II is a flavin containing NADPH: disulfide oxidoreductase (Nocek et al. 2002), and the interactions involving NADP+, FAD, the disulfide, and the 2-ketopropyl-CoM are shown in Figure 7.13b (Clark et al. 2000; Boyd et al. 2006). An analogous mechanism operates in the degradation of epichlorohydrin (1-chloro-2,3-epoxypropane) by the same strain (Small et al. 1995). c. The metabolism of acetone that converges with the metabolism of propene oxides in Xanthobacter sp. strain Py2 is accomplished by an ATP-dependent carboxylase with production of acetoacetate (Sluis et al. 1996, 2002).

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Aliphatic Compounds

307 O

CH 3 C

CH 3

O

H

C

C

H

H C

H

H

CH 3

OH CoM

S

CH 2

OH

CH 3

C

H CoM

S

CH 2

C

H

CH 3 NAD

NAD

CoM

S

CH 2

CO CH3

CO 2

HO2C CH 2 COCH3

FIGURE 7.13

Reaction of epoxide with coenzyme M.

C

CH 3

CH CH 2

FIGURE 7.14

CH 2 OH

CH 2

CH 2 CH 3

C

O

CH 3

CH CH 2

C

SGl

CH CH 2

CH 2 OH CH 3

CH CH CH 2

Glutathione-mediated degradation of isoprene.

5. Rhodococcus sp. strain AD45 carried out the transformation of 2-methyl-1,3-butadiene (isoprene), and both cis- and trans-dichloroethenes to the epoxides (van Hylckama Vlieg et al. 1998). Degradation of the diene (Figure 7.14) takes place by a pathway involving a glutathione S-transferase that is able to react with the epoxides and a conjugate-specific dehydrogenase that produces 2-glutathionyl-2-methylbut-3-enoate (van Hylckama Vlieg et al. 1999, 2000). 6. The degradation of vinyl chloride and ethene has been examined in Mycobacterium sp. strain JS 60 (Coleman and Spain 2003) and in Nocardioides sp. strain JS614 (Mattes et al. 2005). For both substrates, the initially formed epoxides underwent reaction with reduced coenzyme M and, after dehydrogenation and formation of the coenzyme A esters, reductive loss of coenzyme M acetate resulted in the production of S-acetyl-coenzyme A. The reductive fission is formally analogous to that in the glutathione-mediated reaction. Epoxides may be formed from alkenes during degradation by Pseudomonas oleovorans, although octan-1,2-epoxide is not further transformed, and degradation of oct-1-ene takes place by ω-oxidation (May and Abbott 1973; Abbott and Hou 1973). The ω-hydroxylase enzyme is able to carry out either hydroxylation or epoxidation (Ruettinger et al. 1977). Considerable attention has been directed to the epoxidation of alkenes on account of interest in the epoxides as industrial intermediates. The wide metabolic capability of MMO, which has already been noted, has been applied to the epoxidation of C2, C3, and C4 alkenes (Patel et al. 1982). A large number of propane-utilizing bacteria are also effective in carrying out the epoxidation of alkenes (Hou et al. 1983). Especially valuable is the possibility of using microorganisms for resolving racemic mixtures of epoxides. For example, this has been realized for cis- and trans-2,3-epoxypentanes

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Environmental Degradation and Transformation of Organic Chemicals CH 3 O

CH 3

C2 H5

O

H

;

H

H

(2S, 3S)

O

H

C2H5

H

(2S, 3R)

H

O

H

C2 H5

;

CH 3

C 2 H5

(2R, 3S)

FIGURE 7.15

CH 3

H

(2R, 3S)

Biodegradability of enantiomeric of epoxides of cis- and trans-pent-2-enes.

using a Xanthobacter sp., which is able to degrade only one of the pairs of enantiomers leaving the other intact (Figure 7.15) (Weijers et al. 1988). Bacterial epoxidation of alkenes and fungal enzymatic hydrolysis of epoxides have been reviewed in the context of their application to the synthesis of enantiomerically pure epoxides and their derivatives (Archelas and Furstoss 1997). One of the disadvantages of using bacteria, which may carry out undesirable degradation, may sometimes be overcome by the use of fungi (Archelas and Furstoss 1992), although the initially formed epoxides are generally hydrolyzed by fungal epoxide hydrolase activities. Other aspects of epoxide formation and degradation are worth noting, particularly on account of their biotechnological relevance. a. Although Mycobacterium sp. strain E3 is able to degrade ethene via the epoxide, the epoxide-degrading activity is highly specific for epoxyethane and degradation requires reductant generated from glycogen or trehalose storage material (de Haan et al. 1993). b. In Xanthobacter sp. strain Py2 both the alkene monooxygenase and the epoxidase are induced by C2, C3, and C4 alkenes, and also by chlorinated alkenes including vinyl chloride, cis- and trans-dichloroethene, and 1,3-dichloropropene (Ensign 1996). As an alternative to epoxidation of alkenes, hydration is occasionally used, for example in degradation of the ring-fission products from dioxygenation and dehydrogenation of tetralin (Figure 7.16) (Hernáez et al. 2000, 2002). Although the aerobic degradation of squalene involves oxidation and fission at ∆10–11, degradation under anaerobic conditions involves hydration and carboxylation (Rontani et al. 2002) (Figure 7.17). Anaerobic Conditions Degradation of hex-1-ene has been observed in a methanogenic consortium (Schink 1985a) that converted the substrate into methane, and a plausible pathway involving hydration and oxidation was suggested.

ALKYNES The degradation of alkynes has been the subject of sporadic interest during many years, and the pathway has been clearly delineated. It is quite distinct from those used for alkanes and alkenes, and is a reflection of the enhanced nucleophilic character of the alkyne C≡C bond. The initial step is hydration of the triple bond followed by ketonization of the initially formed enol. This reaction operates during the degradation of acetylene itself (de Bont and Peck 1980), acetylene carboxylic acids (Yamada and Jakoby 1959), and more complex alkynes (Figure 7.18) (Van den Tweel and de Bont 1985). It is also appropriate to note that the degradation of acetylene by anaerobic bacteria proceeds by the same pathway (Schink 1985b).

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Aliphatic Compounds

309 OH

OH OH

OH

O HO2C

CO2H OH

OH CO2H CO2H

HO2C OH

FIGURE 7.16

O

CHO + CH3CO−CO2H

HO2C

Aerobic degradation of tetralin by hydration.

CH 2 CH 2

HO CH 2 CH 2

CO 2

O CH 2 CH 2

FIGURE 7.17

Aerobic degradation of squalene.

FIGURE 7.18

Aerobic biodegradation of but-3-ynol.

+

HO 2 C

Acetylene is able to support the growth of the anaerobe Pelobacter acetylenicus (Schink 1985b), and undergoes initial hydration to acetaldehyde followed by dismutation into acetate and ethanol. Although the enzyme is stable in air, it requires a strong reductant such as Ti(III) or dithionite for its activity (Rosner and Schink 1995). The hydratase is a tungsten iron–sulfur protein with a [4Fe–4S] cluster in the reduced state and contains a molybdopterin cofactor conjugated with guanosine monophosphate (Meckenstock et al. 1999).

ALKANOLS Aerobic Conditions Essential aspects of the aerobic degradation of alkanols have already been covered in the context of the degradation of alkanes. Only one additional comment is added: the initial reactions involved

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in the metabolism of higher alkanes (>C1) are formally similar to those used for the metabolism of methane, and the soluble alkanol dehydrogenases also contain PQQ (references in Anthony 1992). Enzymatically, however, the details may be more complex since, for example, a number of distinct alcohol and fatty acid dehydrogenases have been isolated from an Acinetobacter sp. during the metabolism of hexadecane (Singer and Finnerty 1985a,b). Anaerobic Conditions Anaerobic oxidation of methanol—though not incorporation into cellular material—has been observed in a sulfate-reducing bacterium (Braun and Stolp 1985). Ethanol may be converted into methane by Methanogenium organophilum (Widdel 1986; Frimmer and Widdel 1989), and oxidation of primary alkanols has been demonstrated in Acetobacterium carbinolicum (Eichler and Schink 1984). Secondary alcohols such as propan-2-ol and butan-2-ol may be used as hydrogen donors for methanogenesis with concomitant oxidation to the corresponding ketones (Widdel et al. 1988). An NAD-dependent alcohol dehydrogenase has been purified from Desulfovibrio gigas that can oxidize ethanol, and it has been shown that the enzyme does not bear any relation to classical alcohol dehydrogenases (Hensgens et al. 1993). The metabolism of 1,2-diols has attracted considerable attention, and in particular that of glycerol in view of its ubiquity as a component of lipids. Widely different pathways have been found of which two are given as illustration: a. Anaerovibrio glycerini ferments glycerol to propionate (Schauder and Schink 1989) and Desulfovibrio carbinolicus to 3-hydroxypropionate (Nanninga and Gottschal 1987). b. Desulfovibrio alcoholovorans converts glycerol into acetate, and 1,2-propandiol to acetate and propionate (Qatibi et al. 1991).

ALKANONES These may be degraded under aerobic conditions by a sequence involving Baeyer–Villiger monooxygenation to lactones followed by hydrolysis. The resulting alkanols are then degraded to α,ω-carboxylic acids by the reactions following terminal hydroxylation, which have already been described (Figure 7.4). This is used in one of the pathways used for the degradation of 6,10,14-trimethylpentadecan-2-one by Marinobacter sp. strain CAB that involves insertion of an oxygen atom between C2 and C3 and subsequent hydrolysis and oxidation with loss of C2 and C3 (Rontani et al. 1997). The aerobic metabolism of acetone by Xanthobacter sp. strain Py2 involves at ATP-dependent carboxylation (Sluis et al. 1996), with the direct formation of acetoacetate. The carboxylase has been characterized from Xanthobacter autotrophicus strain Py2, and from the anaerobic phototroph Rhodobacter capsulatus strain B10 (Sluis et al. 2002). It has an α 2β2γ2 subunit composition and EPR measurements confirm the presence of Mn2+ (Boyd et al. 2004). The anaerobic degradation of ketones also involves carboxylation followed by hydrolysis (Platen and Schink 1990; Hirschler et al. 1998) (Figure 7.19). There are two subsequent pathways in anaerobes: (1) oxidation to CO2 in denitrifying and sulfate reducing bacteria by the acetyl-CoA-CO dehydrogenase pathway and (2) an anabolic pathway by a modified tricarboxylate-glyoxylate cycle (Janssen and Schink 1995a). Methylglyoxal is a metabolite of unregulated carbohydrate metabolism that is toxic to the cell. Detoxification in Escherichia coli can be achieved by several mechanisms (Ko et al. 2005): (a) by successive reduction to hydroxyacetone and 1,2-propandiol, (b) by reduction to 2-hydroxyacetaldehyde, and (c) by the activity of glyoxalase to d-lactate (Figure 7.20) (Clugston et al. 1998). In Clostridium beijerinckii, however, detoxification is carried out only by reduction to 1,2-propandiol (Liyanage et al. 2001).

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Aliphatic Compounds

311 CO2H

RCH2COCH3

FIGURE 7.19

Anaerobic degradation of ketones. GSH

OH SG

FIGURE 7.20

RCH2CO2 H + CH3CO2H

RCH−COCH3

CH 3 HO

C

C

OH SG

O SG

H CH3

OH C

CO2H

Conversion of methylglyoxal into d-lactate.

ALKYLAMINES AND AMINO ACIDS Aerobic Conditions Alkylamines The initial reaction in the biodegradation of primary alkylamines is conversion into the aldehyde and subsequent reactions converge on those for the degradation of primary alkanes. There are a number of important details in this apparently straightforward reaction: 1. The conversion of alkylamines into aldehydes may be accomplished by different mechanisms: (a) by oxidases that convert molecular oxygen to H2O2 or (b) by dehydrogenases. Which of these is used depends on the organism. For example, the dehydrogenase is used by Pseudomonas aeruginosa ATCC 17933, P. putida ATCC 12633, and the methylotroph Paracoccus versutus ATCC 25364. In contrast, Klebsiella oxytoca ATCC 8724, Escherichia coli ATCC 9637, and Arthrobacter sp. NCIB 11625 use a copper-quinoprotein amine oxidase (Hacisalihoglu et al. 1997) that contains copper at the active site (Wilmot et al. 1999). An inducible primary amine dehydrogenase in a strain of Mycobacterium convolutum had diverse degradative capability with a broad specificity, and was involved in the degradation of 1- and 2-aminopropane and 1,3-diaminopropane. The products were assimilated by the methylmalonate pathway, or by formation of C2 + C1 fragments (Cerniglia and Perry 1975). 2. An alternative pathway for methylamine in Pseudomonas sp. strain MA and Hyphomicrobium is initiated by the formation of N-methylglutamate with the loss of NH4+, and is mediated by an FMN-containing synthetase (Pollock and Hersh 1973). This is dehydrogenated by a flavoprotein to glutamate and formaldehyde (Bamforth and Large 1977), while established pathways metabolize the resulting formaldehyde. 3. The transformation of 2-aminopropane to alaninol by a strain of Pseudomonas sp. KIE171 involves reaction with glutamate in an ATP-dependent reaction to produce a γ-glutamide. Introduction of a hydroxyl group into the methyl group is followed by hydrolysis to regenerate glutamate, and dehydrogenation to alanine (Wäsch et al. 2002). 4. Species of Pseudomonas can utilize the α,ω-diamines putrescine, spermidine, and spermine as sources of carbon and nitrogen, and putrescine can be produced from the secondary amine spermidine together with 1,3-diaminopropane (Dasu et al. 2006). Putrescine (1,4-diaminobutane) is an intermediate in the arginine decarboxylase (ADC) pathway of l-arginine degradation, which is described later, and can be degraded, by two pathways: a. In Pseudomonas aeruginosa via the analogous intermediates used for primary amines— oxidative deamination to 4-aminobutyraldehyde, followed by dehydrogenation to 4-aminobutyrate and oxidative deamination to succinate semialdehyde (Lu et al. 2002) b. In Escherichia coli by an ATP-dependent production of γ-glutamyl putrescine followed by conversion of the terminal–CH2NH2 group into carboxylate and removal of glutamate (Kurihara et al. 2005). Succinate semialdehyde is then produced by reactions, which have already been outlined.

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The degradation of secondary or tertiary amines has been examined most extensively with dimethylamine and trimethylamine. The latter can be degraded by two mechanisms—dehydrogenation to dimethylamine and formaldehyde or monooxygenation to the N-oxide followed by demethylation. The resulting dimethylamine can then be degraded by dehydrogenation. Species of Paracoccus able to degrade tetramethylammonium hydroxide have been assigned to the species Pa. kocurii (Ohara et al. 1990), and those degrading N,N′-dimethylformamide to Pa. aminophilus and Pa. aminovorans (Urakami et al. 1990). Although Aminobacter aminovorans (Pseudomonas aminovorans) was able to utilize methylamine and trimethylamine, it was unable to use methane, methanol, or dimethylamine (Urakami et al. 1992). Considerable effort has been directed to the degradation of complexing agents. The degradation of nitrilotriacetate (NTA), ethylenediamine tetraacetate (EDTA), and related compounds display some important features: a. The degradation of NTA takes place by successive loss of glyoxylate (Cripps and Noble 1973; Firestone and Tiedje 1978), and the monooxygenase system consists of two components, both of which are necessary for hydroxylation (Uetz et al. 1992); b. The first step in the degradation of EDTA is carried out by a flavin-dependent monooxygenase, which has been purified (Witschel et al. 1997; Payne et al. 1998), and results in the formation of glyoxal and N,N-ethylenediaminediacetate (EDDA). The ultimate formation of ethylenediamine and glyoxylate is, however, carried out by an oxidase (Liu et al. 2001). It may be assumed that an analogous pathway exists for the degradation of triethylenediaminepentaaacetic acid (DTPA); c. Iminodisuccinate, which is a potential replacement for ETDA, has two asymmetric centers and, since the C−N lyase that cleaves it to d-aspartate and fumarate is stereospecific, degradation is initiated by the activity of epimerases to the R,S-enantiomer. d-Aspartate is then isomerized to l-aspartate that undergoes elimination to fumarate (Figure 7.21) (Cokesa et al. 2004). The degradation of morpholine has been described in several strains of Mycobacterium, including M. chelonae (Swain et al. 1991), M. aurum strain MO1 (Combourieu et al. 1998), and Mycobacterium strain RP1 (Poupin et al. 1998). Initiation of the degradation of the cyclic secondary amines morpholine, piperidine, and pyrrolidine by the strain of Mycobacterium RP-1 is carried out by cytochrome P450 hydroxylation (Poupin et al. 1998). The first step in the degradation of the tertiary amine nicotine by Arthrobacter nicotinovorans involves hydroxylation to 6-hydroxynicotine, which is followed by fission of the N-methylpyrrolidine ring by an oxidase (Dai et al. 1968) and is discussed in Part 1 of Chapter 10. Amino acids A number of factors complicate the aerobic metabolism of amino acids—different enzymes may be used even for the same amino acid; the enzymes may be inducible or constitutive depending on their function; α-ketoacids may be produced by deamination or amines by decarboxylation. 1. Pyridoxal 5′phosphate enzymes are generally used when amino acids are supplied as sources of carbon or nitrogen. An aldimine is produced initially and is isomerized to

HO2C−CH2 HO 2 C

HH N H

CH 2 CO 2 H

HO 2 C

H +

CO 2 H

H2 N H

H

CO 2 H

CH 2

CO 2 H

CH 2 CO 2 H

HO 2 C

H

H CO 2H

D-asp

H2N

CO 2H

H

CO 2 H

L-asp

FIGURE 7.21 Aerobic degradation of iminodisuccinate.

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Aliphatic Compounds

313

the corresponding ketimine through loss of the α-proton that is facilitated by a lysine group on the enzyme. The ketimine is then hydrolyzed to pyridoxamine phosphate and an α-ketocarboxylic acid that is incorporated into established catabolic pathways. There are some additional important features: a. Lysine. In pseudomonads, there are different pathways for the degradation of l-lysine (Figure 7.22a) and d-lysine (Figure 7.22b) (Muramatsu et al. 2005). Genetic evidence suggests their connection in Pseudomonas putida strain KT2440 (Revelles et al. 2005), which is an efficient colonizer of important commercial plants in the rhizosphere (Molina et al. 2000). b. L- and D -Leucines are degraded by Pseudomonas aeruginosa in a pathway analogous to that for branched-chain alkanes. This involves formation of isovaleryl-CoA, dehydrogenation to methylcrotonly-CoA that is degraded by terminal carboxylation to 3-methylgutaconyl-CoA before hydroxylation, and fission to acetyl-CoA and acetoacetate (Figure 7.23a) (Massey et al. 1974; Höschle et al. 2005). c. 1-Aminocyclopropane-1-carboxylate. This somewhat unusual amino acid is the immediate precursor of the important plant hormone ethene, and the bacterial deaminase in the rhizosphere has therefore attracted considerable attention (Belimov et al. 2001). Unlike other amino acids, 1-aminocyclopropane-1-carboxylate does not contain the α-proton that is required for conventional deamination, and its degradation involves concomitant fission of the cyclopropane ring to α-ketobutyrate (Figure 7.23b). The mechanism has been examined and several possibilities have been put forward that are consistent with the x-ray structure of the enzyme (Karthikeyan et al. 2004) and with isotopic evidence (Walsh et al. 1981). d. L-Arginine. The degradation of arginine can be accomplished by no less than four different mechanisms depending on the circumstances (Nakada and Itoh 2002). The deiminase pathway enables organisms to grow anaerobically with l-arginine, and includes Bacillus licheniformis, Halobacterium halobium, Streptococcus faecalis, and to a limited extent Pseudomonas aeruginosa. The four pathways are summarized:

CH 2 NH 2 (a)

CH 2 NH 2

(CH 2 )3 CH

(CH 2 )3

NH 2

CONH 2

CH 2 NH 2

CO 2 H

(CH 2 )3

(CH 2 )3

CO 2 H

CO 2 H

CO 2 H L-lysine

CH 2 NH 2 (b) H2 N

(CH 2 )3

CH 2 NH 2 (CH 2 )3

CH

CO

CO 2 H

CO 2 H

N

CO 2 H

N H

CO 2 H

N

CO 2 H

D-lysine

CO 2 H

CO 2 H

CO 2 H

(CH 2 )3

(CH 2 )3

(CH 2 )3

CH

NH 2

CO 2 H

FIGURE 7.22

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CO

CO 2 H

CO 2 H

Degradation of (a) l-lysine and (b) d-lysine.

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CO 2 H

CH 3

CH 3

CH 2 O

CH 3

CH 3

OH

HO2C−CH2−C−CH2−COSCoA

CH 3

CH−CH2−CO−CO2H

CO 2 H

CH−COSCoA

CH 3

CH 3

CH 3

CH 3

CH

CH−COSCoA

HO2C−CH2−CO−CH3 + CH 3 COSCoA

CH−CH2−COSCoA

Degradation of (a) d- and l-leucine, (b) 1-aminocyclopropane-1-carboxylate.

CO 2 H

NH 2

HO2C−CH2−C

CH 3

NH 2

CH 3

FIGURE 7.23

(b)

(a)

314 Environmental Degradation and Transformation of Organic Chemicals

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Aliphatic Compounds

315

Arginine deiminase (ADI) pathway (Mercenier et al. 1980); arginine → citrulline → ornithine + carbamoyl phosphate (ADC) pathway (Lu et al. 2002); arginine → agmatine → N-carbamoylputrescine → putrescine → 4-aminobutyrate Arginine dehydrogenase (ADH) pathway (Nakada and Itoh 2002); 4-aminobutyrate + urea

arginine → 2-ketoarginine →

Arginine succinyltransferase (AST) pathway (Schneider et al. 1998); arginine → N 2 succinylarginine → N 2-succinylornithine → N 2-succinylglutamate → glutamate 2. As an alternative to oxidative deamination, decarboxylation is involved in polyamine biosynthesis and in the protection of enteric bacteria from acidic conditions. Under these conditions, the electrons used to form the ketimine are provided by decarboxylation in place of those from the α-methine group. a. In Escherichia coli—and probably other Enterobacteriaceae—amino acids play a role in the resistance to extremely acid conditions through removal of intracellular protons by decarboxylation. Tolerance is mediated by decarboxylase antiporter systems including arginine/agmatine, glutamate/γ-aminobutyrate, and lysine/cadaverine (Iyer et al. 2003). The first of these also is involved in the acid resistance of Salmonella enterica serovar Typhimurium under anaerobic conditions (Kieboom and Abee 2006). b. Polyamines have an important function in contributing to the stability of molecular complexes in the cell. The biosynthesis of putrescine generally involves the decarboxylation of arginine or ornithine. There are two decarboxylases for arginine and ornithine, both of which are dependent on pyridoxal 5′-phosphate and Mg2+ (references in Tabor and Tabor 1985), and probably also for lysine (Lemonnier and Lane 1998; Kikuchi et al. 1997). The synthesis of these alternative enzymes is controlled by the conditions under which the cells are grown—aeration at neutral pH, or semianaerobic at low pH. Exceptionally a pyruvoyl-dependent ADC in Methanococcus jannaschii functions in polyamine synthesis (Graham et al. 2002). The 3-aminopropyl group in spermine and spermidine is produced by decarboxylation of S-adenosylmethionine, which is explained in the next paragraph. 3. A completely distinct enzyme has been found in a number of organisms, which carry out the metabolism of amino acids. In this group, a pyruvoyl group is covalently bound to the active enzyme that is produced from a proenzyme in a self-maturation process (Toms et al. 2004). The proenzyme contains a serine residue that undergoes rearrangement to an ester followed by conversion into the β-chain of the enzyme and a dehydroalanine residue that forms the N-terminal pyruvoyl group of the α-chain. This type of enzyme has been found for a number of important decarboxylations: a. Histidine decarboxylation in Lactobacillus including L. buchneri (Huynh and Snell 1985; Martín et al. 2005), which is the most extensively studied. b. Arginine decarboxylation in Methanococcus jannaschii (Graham et al. 2002). c. S-adenosylmethionine carboxylase in Escherichia coli (Markham et al. 1982; Anton and Kutny 1987), Sulfolobus solfataricus (Cacciapuoti et al. 1991), M. jannaschii (Lu and Markham 2004), and Thermotoga maritima (Toms et al. 2004). S-adenosylmethionine carboxylase is the source of the propylamine in the polyamines spermine and spermidine. The activity of spermine synthase introduces this into spermidine and spermine, which has already been noted. It is worth pointing out that, whereas the inducible histidine decarboxylase

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Environmental Degradation and Transformation of Organic Chemicals

from Gram-positive bacteria belongs to the pyruvoyl group, those from the Gram-negative Enterobacteriaceae including Klebsiella planticola and Enterobacter aerogenes are pyridoxal 5′-phosphate-dependent (Kamath et al. 1991). Anaerobic Conditions A range of C1 compounds such as methylamine, trimethylamine, and betaine can be used by methanogens for the synthesis of methane. Anaerobic degradation of some amino acids by clostridia involves rearrangements and dismutation between pairs, and the biogenic formation of toluene from phenylalanine and of p-cresol from tyrosine is worth noting. Few investigations have been devoted to the anaerobic degradation of alkylamines, though there has been considerable interest in the metabolism of trimethylamine, choline, and glycine betaine as a source of methane in marine sediments (King 1984). Some aspects of the metabolism of alkylamines are used as illustration: 1. Methane formation from trimethylamine by Methanosarcina barkeri has been demonstrated (Hippe et al. 1979; Patterson and Hespell 1979), and the metabolically versatile organisms Methanococcoides methylutens (Sowers and Ferry 1983) and Methanolobus tindarius (Konig and Stetter 1982), which use methylamines and methanol for methane formation, have been described. The methanogen Methanomethylovorans hollandica is able to use a range of methyl substrates including mono-, di-, and trimethylamine as well as dimethyl sulfide and methanethiol (Lomans et al. 1999). 2. The metabolism of betaine, which is an important osmoregulatory solute in many organisms, has been studied and different metabolic pathways have been revealed. a. Demethylation to dimethylglycine by Eubacterium limosum (Müller et al. 1981) and by strains of Desulfobacterium sp. (Heijthuijsen and Hansen 1989a). b. Fermentation by a strain of Desulfuromonas acetoxidans with the production of trimethylamine and acetate (Heijthuijsen and Hansen 1989b), while the same products are formed by Clostridium sporogenes in a Stickland reaction with alanine, valine, leucine, or isoleucine (Naumann et al. 1983). 3. The degradation of amino acids has been examined in detail, particularly in clostridia (Barker 1981), and these investigations have revealed a number of important reactions not encountered in other degradations. a. Coenzyme B12–mediated rearrangements have been elucidated as an important reaction in the degradation of glutamate and ornithine, and subsequently, led to a detailed investigation into the role of this rearrangement in other reactions (Barker 1972). An outline of the β-methylaspartate pathway for the degradation of glutamate by Clostridium tetanomorphum is given in Figure 7.24. The involvement of radicals in a number of rearrangements carried out by clostridia during fermentation of amino acids has been discussed in a review (Buckel and Golding 1999). These include S-lysine to CH 3−CH(CO2H)−CH(NH 2)−CO2H

HO 2C−CH 2−CH2−CH(NH 2)−CO2H

HO 2C

H

OH HO 2C−C−CH 2CO 2H

CH 3

CH 3−CO2H

+

CH3−CO−CO2H

CO 2H CH 3

FIGURE 7.24

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Biodegradation of glutamate by Clostridium tetanomorphum.

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Aliphatic Compounds CH 2−CH(NH2)CO2H

FIGURE 7.25

N H

CH2−CO−CO2H

CH2−CO2H

CH3

Metabolism of phenylalanine.

CO2H

FIGURE 7.26

317

+

CO2H NH2

+

+

CO2 + NH3

Stickland reaction between l-proline and l-alanine.

S-β-lysine (lysine 2,3-aminomutase), S-glutamate to (2S,3S)-3-methylaspartate (glutamate mutase), and 2-methyleneglutarate—produced during the fermentation of nicotinate by Clostridium barkeri—to (R)-3-methylitaconate (2-methyleneglutarate mutase). b. The transformation of aromatic amino acids has been examined, and it has been shown (Elsden et al. 1976; D′Ari and Barker 1985) that in Cl. difficile these compounds are possible biogenic sources of 4-methyl phenol. In an organism designated “Cl. aerofoetidum” toluene may be produced from phenylalanine (Pons et al. 1984) and the postulated reactions are schematically shown in Figure 7.25. c. An unusual dismutation reaction involving pairs of amino acids has been studied extensively in Cl. sporogenes (Stickland 1934, 1935a,b). The reaction can be carried out by many other clostridia and has been summarized (Barker 1961). The products from proline and alanine (Stickland 1935a) illustrate the reaction and are shown in Figure 7.26, and the mechanism whereby glycine is reduced to acetate by glycine reductase from Cl. sticklandii has been elucidated (Arkowitz and Abeles 1989). The involvement of betaine in the reaction has already been noted.

ALKANOIC ACIDS Aerobic Conditions Oxalate and Glyoxylate Oxalate is synthesized from acetyl-CoA and oxalacetate in Burkholderia glumae (Li et al. 1999) and in the wood-rotting basidiomycete Fomitopsis palustris (Munir et al. 2001). Oxalate is also produced in a few degradations, although details of its metabolism are not fully resolved. Oxalate decarboxylase produces formate and CO2, and although its physiological significance in bacteria is not fully resolved, the enzyme in Bacillus subtilis requires manganese for activity (Tanner et al. 2001), and is induced not by oxalate, but by an acidic pH (Tanner and Bornemann 2000). Glyoxylate is produced during several aerobic degradations: a. 1,2-Dichloroethane by both the hydrolytic (Janssen et al. 1985) and oxidative (Hage and Hartmans 1999) pathways b. Allantoate that is a product of purine metabolism (Cusa et al. 1999) c. Morpholine by Mycobacterium aurum strain MO1 (Combourieu et al. 1998) d. Cyclic ethers by the fungus Cordyceps sinensis (Nakamiya et al. 2005). The pathway for bacterial degradation of glyoxylate involves condensation to tartronic semialdehyde by glyoxylate carboligase, reduction to glycerate, and elimination of 3-phosphoglycerate to produce pyruvate (Ornston and Ornston 1969) (Figure 7.27). In the brown-rot fungus Tyromyces

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Environmental Degradation and Transformation of Organic Chemicals O CHO C

HO 2 CCH(OH)CHO

HO 2 CCH(OH)CH 2 OH

CH 3 COCO 2 H

OH O CH CO 2 H

FIGURE 7.27

Aerobic degradation of glyoxylate.

palustris, however, glyoxylate is dehydrogenated to oxalate by a cytochrome c–dependent enzyme (Tokimatsu et al. 1998). Anaerobic Conditions Oxalate, Malonate, and Glycolate Although these are rather unusual substrates, their degradation by anaerobic bacteria has been described. 1. Oxalobacter vibrioformis uses oxalate as sole source of energy using acetate as principal carbon source (Dehning and Schink 1989a). In O. formigenes, oxalate is metabolized to formate and CO2 by the coupled reaction of formyl-CoA transferase and a thiamin diphosphate–dependent oxalyl-CoA decarboxylase, which is activated by adenosine diphosphate (Berthold et al. 2005). Oxalate and formate form a one-to-one antiport system, which involves the consumption of an internal proton during decarboxylation, and serves as an indirect proton pump (Ruan et al. 1992) to generate ATP by decarboxylative phosphorylation (Anantharam et al. 1989). Cell-free extracts contain the enzymes of the biosynthetic glycerate pathway via glyoxylate and tartronic semialdehyde (Cornick and Allison 1996). Oxalyl-CoA decarboxylase has been characterized from several species of Bifidobacterium including B. lactis, B. animalis, B. longum, and B. adolescens (Federici et al. 2004), and it was suggested that such strains could be important in the intestinal degradation of oxalate. 2. Malonomonas rubra is a microaerotolerant fermenting organism that decarboxylates malonate to acetate (Dehning and Schink 1989b). The organism contains high concentrations of c-type cytochromes that are not involved in the metabolism of the substrate, and are presumably remnants of the sulfur-reducing relatives of the organism (Kolb et al. 1998). Biotin-dependent carboxylases couple the decarboxylation to the transport of Na+ across the cytoplasmic membrane and use the electrochemical potential ∆µ Na+ to mediate the synthesis of ATP (Dimroth and Hilbi 1997). 3. Glycolate or fumarate is fermented by an organism belonging to the family Lachnospiraceae to acetate, succinate, and CO2 without the formation of hydrogen (Janssen and Hugenholtz 2003). Higher Carboxylic Acids The aerobic degradation of these has been covered in the discussion on alkanes and need not be repeated here. Anaerobic Conditions The anaerobic degradation of aliphatic carboxylic acids is of great ecological importance, since compounds such as acetate, propionate, or butyrate may be the terminal fermentation products of organisms degrading more complex compounds including carbohydrates, proteins, and lipids, while long-chain acids are produced by the hydrolysis of lipids (Zeikus 1980; Mackie et al. 1991). Degradation of aliphatic carboxylic acids by sulfate-reducing bacteria was traditionally restricted to

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319

lactate and its near relative, pyruvate, but recent developments have radically altered the situation and increased the spectrum of compounds which can be oxidized to CO2 at the expense of sulfate reduction. In the following paragraphs, an attempt will be made to present a brief summary of the anaerobic degradation of the main groups of aliphatic compounds. Studies on the anaerobic degradation of alkanoic acids have been carried out using both pure cultures and syntrophic associations. Acetate For the degradation of acetate, two different reactions may take place: oxidation to CO2 or dismutation to methane and CO2. To some extent, as it will emerge, segments of both pathways are at least formally similar, although the mechanisms for anaerobic degradation of these apparently simple compounds are quite subtle. Degradation of acetate and butyrate can be accomplished by Desulfotomaculum acetoxidans (Widdel and Pfennig 1981) and of acetate by Desulfuromonas acetoxidans (Pfennig and Biebl 1976) and species of Desulfobacter (Widdel 1987), while propionate is degraded by species of Desulfobulbus (Widdel and Pfennig 1982; Samain et al. 1984). A sulfur-reducing organism with a much wider degradative capability than Desulfuromonas acetoxidans has been isolated (Finster and Bak 1993), and this organism is capable of accomplishing the complete oxidation of, for example, propionate, valerate, and succinate. The oxidation of acetate under anaerobic conditions can take place by different pathways, both of which have been investigated in detail and their enzymology delineated (Thauer et al. 1989). 1. Oxidation may take place by a modified tricarboxylic acid cycle in which the production of CO2 is coupled to the synthesis of NADPH and reduced ferredoxin, and the dehydrogenation of succinate to fumarate is coupled to the synthesis of reduced menaquinone. This pathway is used, for example, by Desulfuromonas acetoxidans and in modified form by Desulfobacter postgatei. 2. In contrast, dissimilation of acetate may take place by reversal of the pathway used by organisms such as Clostridium thermoaceticum for the synthesis of acetate from CO2. In the degradation of acetate, the pathway involves a dismutation in which the methyl group is successively oxidized via methyl THF to CO2 while the carbonyl group is oxidized via bound carbon monoxide. Such THF-mediated reactions are of great importance in the anaerobic degradation of purines, which is discussed in Chapter 10, Part 1. Acetate may also be converted into methane by a few methanogens belonging to the genus Methanosarcina. The methyl group is initially converted into methyltetrahydromethanopterin (corresponding to methyltetrahydrofolate in the acetate oxidations discussed above) before reduction to methane via methyl-coenzyme M; the carbonyl group of acetate is oxidized via bound CO to CO2. Propionate, Succinate, and Long-Chain Carboxylic Acids Degradation of propionate is an intermediate step in the decomposition of organic matter in anaerobic environments, and growth at the thermodynamic limit is facilitated by syntrophy with methanogenic bacteria. For example, the syntrophic population of anaerobic organisms that carried out the oxidation of propionate in a rice field was examined and showed the presence of the genera Syntrophobacter, Smithella, and Pelotomaculum. Among the archaea, species of Methanobacterium, Methanosarcina, and members of the Rice Cluster I were found (Lueders et al. 2004). A number of important aspects are noted: 1. Both the synthesis of propionate and its metabolism may take place under anaerobic conditions. In Desulfobulbus propionicum, degradation could plausibly take place by reversal of the steps used for its synthesis from acetate (Stams et al. 1984)—carboxylation of propionate to methylmalonate followed by coenzyme B12–mediated rearrangement to succinate, which then enters the tricarboxylic acid cycle. The converse decarboxylation of succinate to propionate has been observed in Propionigenium modestum (Schink and Pfennig 1982),

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320

Environmental Degradation and Transformation of Organic Chemicals 2CH3CH2CO2H

CH3CH2COCHCO2H

CH3CH2COCH2CH2CO2H

CH3CH2 CH CH

CH 2 CO 2 H

CH 3 CH3CH2CH2CH(OH)CH2CO2H

CH3CH2CH2COCH2CO2H

CH3CH2CH2CO2H + CH3CO2H

FIGURE 7.28 Anaerobic degradation of propionate by Smithella propionica.

Sporomusa malonica (Dehning et al. 1989), Selenomonas acidaminovorans (Guangsheng et al. 1992), and Propionigenium maris (Janssen and Liesack 1995). 2. Growth of syntrophic propionate-oxidizing bacteria in the absence of methanogens has been accomplished using fumarate as the sole substrate (Plugge et al. 1993). Fumarate plays a central role in the metabolism, since it is produced from propionate via methylmalonate and succinate, while fumarate itself is metabolized by the acetyl-CoA cleavage pathway via malate, oxalacetate, and pyruvate. 3. The pathways and mechanism of interspecies transfer have been examined in syntrophic propionate-oxidizing organisms. a. Two pathways for propionate degradation have been proposed. In Pelotomaculum thermopropionicum the methylmalonyl-CoA pathway is used, and stringent regulation of fumarase is exerted (Kosaka et al. 2006). In Smithella propionica, which produces acetate and butyrate, however, the pathway is initiated by condensation of two molecules of propionate. This is followed by rearrangement—putatively by a coenzyme B12–dependent mutase—to 3-ketohexanoate, and fission to butyrate and acetate (Figure 7.28) (de Bok et al. 2001). b. Although it has been generally assumed that interspecies transfer in propionateoxidizing syntrophs involved H2, there is increasing evidence that formate transfer is actually involved. Components of the syntrophic culture of Syntrophobacter fumaroxidans and Methanospirillum hungatei that degrade propionate were separated by Percoll gradient centrifugation. Levels of formate dehydrogenase, hydrogenase, and formatehydrogen lyase in the components were examined, and it was showed that interspecies electron transfer was carried out primarily via formate (de Bok et al. 2002). The degradation of long-chain carboxylic acids is important in the anaerobic metabolism of lipids and an extensive compilation of the organisms that can accomplish this has been given (Mackie et al. 1991). This capability has been demonstrated in syntrophic bacteria in the presence of hydrogen-utilizing bacteria; for example, β-oxidation of C4 to C8, C5, and C7 carboxylic acids was carried out by the Syntrophomonas wolfei association (McInerney et al. 1981), and of C4 to C10, and C5 to C11 by the Clostridium bryantii syntroph (Stieb and Schink 1985). Acetate and propionate were the respective terminal products from the even- and odd-numbered acids. Single cultures of many sulfate-reducing bacteria are also able to carry out analogous reactions (Mackie et al. 1991). For example, Desulfobacterium cetonicum degrades butyrate to acetate by a typical β-oxidation pathway (Janssen and Schink 1995b). Long-Chain Dicarboxylic Acids Long-chain dicarboxylic acids are produced by a number of reactions: a. The C12 dicarboxylate by α,ω-oxidation of dodecane (Broadway et al. 1993) b. The C7 dicarboxylate (pimelate) as an intermediate in the anaerobic degradation of benzoate (Harwood et al. 1999) c. As intermediates in the degradation of cycloalkanones (Kostichka et al. 2001) d. Atmospheric oxidation of unsaturated fatty acids (Stephanou and Stratigakis 1993). The degradation of pimelate is initiated by formation of the coenzyme A ester and is followed by a series of steps with the production of glutaryl-CoA that is decarboxylated to crotonyl-CoA

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Aliphatic Compounds

321

(Härtel et al. 1993; Gallus and Schink 1994; Harrison and Harwood 2005). This produces two molecules of acetate or, in strictly fermenting organisms butyrate and isobutyrate (Matties and Schink 1992). The genetics of the degradation has been explored (Parke et al. 2001) in Acinetobacter sp. strain ADP1 that is able to degrade dicarboxylic acids up to C14, and the enzymes for the degradation of pimelate in Rhodopseudomonas palustris are contained in an operon in which the PimA ligase that catalyzes the first step is able to accept an unusually wide range of dicarboxylic acids as substrates (Harrison and Harwood 2005). In summary, the anaerobic degradation of alkanoic acids may truly be described as ubiquitous and is carried out by organisms with widely different taxonomic affinity both in pure culture and in syntrophic associations.

AMIDES AND RELATED COMPOUNDS Amides Aromatic amides, carbamates, and ureas are components of a number of important agrochemicals, and the first step in their biodegradation is mediated by the activities of amidases, ureases, and carbamylases with the production of amines (Figures 7.29a through c). The chloroanilines that are formed from many of them as initial products may, however, be substantially more resistant to further degradation. Application of assays for amidase activity, particularly for pyrazinamidase, has been widely used in the classification of mycobacteria (Wayne et al. 1991). Sequential hydrolysis of nitriles to amides and carboxylic acids is well established both in aliphatic (Miller and Gray 1982; Nawaz et al. 1992) and aromatic compounds (Harper 1977; McBride et al. 1986). Degradation of the herbicide bromoxynil may, however, take place by the elimination of cyanide from the ring with the initial formation of 2,6-dibromohydroquinone (Topp et al. 1992) (Figure 7.30). There may be a high

O C2H5O (a)

P

O

C2H5O

NO 2

C2H5O

NO 2

O

P

C2H5O

O (C2H5O)P−OH +

HO

NO 2

NH−CO−CH 2−CH3

NH 2 +

(b) Cl

Cl

Cl

Cl

Cl

NH−CO−N

CH 3

NH 2

O−CH 3

(c) Cl

CH 3−CH2−CO2H

+

CH 3−NH−OCH3

Cl Cl

FIGURE 7.29 Hydrolysis of (a) acetanilides, (b) phenylureas, and (c) carbamates.

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Environmental Degradation and Transformation of Organic Chemicals CN

Br

OH

Br OH

FIGURE 7.30

Br

Br OH

Metabolism of bromoxynil.

degree of specificity in the action of these nitrilases and this may have considerable significance in biotechnology. Some examples are given as illustrations: 1. Racemic 2-(4′-isobutylphenyl)propionitrile is converted by a strain of Acinetobacter sp. into S-(+)-2-(4′-isobutylphenyl)propionic acid with an optical purity > 95% (Yamamoto et al. 1990). 2. The nitrilase from a number of strains of Pseudomonas sp. mediated an enantiomerically selective hydrolysis of racemic O-acetylmandelonitrile to d-acetylmandelic acid R-(−)acetylmandelic acid (Layh et al. 1992). 3. The amidase from Rhodococcus erythopolis strain MP50 was used to selectively convert racemic 2-phenylpropionamide into S-2-phenylpropiohydroxamate. This was converted into the isocyanate by Lossen rearrangement and then by hydrolysis to S-(−)-phenylethylamine (Hirrlinger and Stolz 1997). 4. An amidase from Ochrobactrum anthropi strain NCIMB 40321 has a wide substrate versatility for l-amides, primarily those with an α-amino group (Sonke et al. 2005), while the condensation product of urea and formaldehyde H2N–[CONH–CH2NH]n –CONH2 is hydrolyzed by another strain of O. anthropi (Jahns et al. 1997).

NITRILES There are two pathways for the degradation of nitriles: (a) direct formation of carboxylic acids by the activity of a nitrilase, for example, in Bacillus sp. strain OxB-1 and P. syringae B728a (b) hydration to amides followed by hydrolysis, for example, in P. chlororaphis (Oinuma et al. 2003). The monomer acrylonitrile occurs in wastewater from the production of polyacrylonitrile (PAN), and is hydrolyzed by bacteria to acrylate by the combined activity of a nitrilase (hydratase) and an amidase. Acrylate is then degraded by hydration to either lactate or β-hydroxypropionate. The nitrilase or amidase is also capable of hydrolyzing the nitrile group in a number of other nitriles (Robertson et al. 2004) including PAN (Tauber et al. 2000).

ISONITRILES Isonitriles are widespread metabolites in marine organisms, although little attention has been paid to their degradation. Pseudomonas putida strain N-19-2 can hydrolyze isonitriles to N-substituted formamides (Goda et al. 2001), and the enzyme is active toward cyclohexyl isonitrile and benzyl isonitrile.

SULFONYLUREAS AND THIOCARBAMATES 1. Sulfonylureas are the basis of a large group of herbicides. Cytochrome P450 enzymes in Streptomyces griseolus transform the sulfonylureas by hydroxylation (Omer et al. 1990) leaving the –SO2NHCONH– part of the structure unaltered (Harder et al. 1991). 2. Substantial effort has been given to the degradation of thiocarbamates and the expression of the thcB gene encoding cytochrome P450 in strains of Rhodococcus is expressed constitutively (Shao and Behki 1996). The degradation of S-ethyl N,N′-dipropylthiocarbamate

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Aliphatic Compounds

323

CH 3 CH 2 CH 2

CH 3 CH 2 CH 2 N−CO−S−CH 2CH 3

CH 3 CH 2 CH 2

N−CO−S−CH 2CH 3 CH 3 CH 2 CH OH

CH 3 CH 2 CHO + CH 3CH 2CH 2−NH−CO−S−CH 2CH 3 CH 3 CH 2 CH 2 NH 2 + HO 2C−S−CH 2CH 3

FIGURE 7.31

HO 2C−S−CH 3

CO 2 + CH 3 SH

Aerobic degradation of S-ethyl N,N′-dipropylthiocarbamate.

(EPTC) has been examined in a number of rhodococci and in one of them (strainTE1 given as a species of Arthrobacter sp.) degradation is associated with a 50.5 mDa plasmid (Tam et al. 1987). The principal pathway in Rhodococcus sp. strain JE1 involves hydroxylation of the propyl group followed by loss of propionaldehyde and degradation of N-depropyl EPTC to S-ethylformate and propylamine (Figure 7.31) (Dick et al. 1990). Two of the genes induced by EPTC encode a cytochrome P450 system that carries out the initial hydroxylation with formation of the N-depropylated product and an aldehyde dehydrogenase that converts the aldehyde produced into the corresponding carboxylic acid (Nagy et al. 1995). In addition to these mechanisms, the degradation of thiocarbamates may be carried out in Rhodococcus erythropolis NI86/21 by a herbicide-inducible nonheme haloperoxidase (de Schrijver et al. 1997). Although it is moot whether carbon monoxide, cyanide, and thiocyanate should be included as organic compounds, brief comments on their biodegradation are summarized.

CARBON MONOXIDE Carbon monoxide is produced biotically from hemin and amino acids and in soil, and during fossil fuel incomplete combustion. Although normally toxic to aerobic organisms, microorganisms have been isolated that are able to degrade it, and in a bound form it is an important intermediate in anaerobic reactions. Some aerobic bacteria are able to use carbon monoxide as a source of energy and cell carbon (Meyer and Schlegel 1993), and mycobacteria are able to utilize both carbon monoxide and methanol (Park et al. 2003). CO2 is formed by a carbon monoxide dehydrogenase and is incorporated into biosynthetic reactions by the ribulosebisphosphate pathway (Park et al. 2003). Carbon monoxide is also an important metabolite in reactions carried out by anaerobic bacteria including the degradation of acetate by a pathway that involves a dismutation in which the methyl group is successively oxidized via methyltetrahydrofolate to CO2, while the carbonyl group is oxidized via bound carbon monoxide. Acetate may also be converted into methane by a few methanogens belonging to the genus Methanosarcina. The methyl group is initially converted into methyltetrahydromethanopterin (corresponding to methyltetrahydrofolate in the acetate oxidations) before reduction to methane via methyl coenzyme M; the carbonyl group of acetate is oxidized via bound CO to CO2.

CYANIDE Cyanide is a component of plant cyanogenic glycosides such as amygdalin and linamarin, while a range of structurally diverse diterpenoid isonitriles is produced by marina biota. Cyanide is a toxic waste from the mining and electroplating industries and, although normally toxic to aerobic organisms, microorganisms have been isolated that are able to degrade it. Two mechanisms have emerged for the degradation of cyanide (Fernandez et al. 2004) and the corresponding nitriles: a. Hydrolysis by nitrilase to the carboxylic acid and ammonia b. Hydratases that catalyze the formation of the amides

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Environmental Degradation and Transformation of Organic Chemicals

The mechanism of the first reaction in Pseudomonas fluorescens NCIMB 11764 is complex, and involves both monooxygenation and hydrolysis. Oxygenation is carried out by reduced pterins and is analogous to the hydroxylation of phenylalanine to tyrosine. The overall reaction involves monooxygenation to cyanate followed by reduction to formamide and hydrolysis to formate and NH4+. An alternative degradation for cyanate has, however, been shown in Escherichia coli when an inducible cyanase brings about a recycling reaction between bicarbonate and cyanate to form formamide that is hydrolyzed to CO2 and NH4+ (Johnson and Anderson 1987). A strain of Pseudomonas pseudoalcaligenes isolated by enrichment with cyanide as sole source of nitrogen was able to utilize, in addition, ferrocyanide, nitroferricyanide, and cuprocyanide, and may involve cyanate (Luque-Almagro et al. 2005). Noncyanogenic fungi can degrade cyanide to formamide followed by hydrolysis by a hydratase to formate and ammonia (Dumestre et al. 1997). This pathway is also used by some bacteria (Jandyala et al. 2003). Bacteria also use a number of reactions for the detoxification of cyanide, including monooxygenation to CO2 and ammonia (Wang et al. 1996).

THIOCYANATE Although the degradation of thiocyanate is accomplished by a few bacteria, the chemolithoautrophic Thiobacillus thioparus strain THI 115 is unusual in being able to use this as an energy source. The hydrolase that converts the substrate into NH4+ and COS has been purified, and consists of three subunits (Katayama et al. 1998). The hydrolase showed a high homology to nitrile hydrolase, which suggests a close similarity in the pathways for their degradation—hydrolysis to the amines followed by deamination.

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Swaving J, JAM de Bont, A Westphal, A de Kok (1996) A novel type of pyridine nucleotide-disulfide oxidoreductase is essential for NAD+- and NADPH-dependent degradation of epoxyalkanes by Xanthobacter strain Py2. J Bacteriol 178: 6644–6646. Tabor CW, H Tabor (1985) Polyamines in microorganisms. Microbiol Revs 49: 81–99. Tam AC, RM Behki, SU Khan (1987) Isolation and characterization of an S-ethyl-N,N-dipropylthiocarbamatedegrading Arthrobacter strain and evidence for plasmid-associated S-ethyl-N,N-dipropylthiocarbamate degradation. Appl Environ Microbiol 53: 1088–1093. Tanaka A, M Ueda (1993) Assimilation of alkanes by yeasts: functions and biogenesis of peroxisomes. Mycol Res 98: 1025–1044. Tanner A, L Bowater, SA Fairhurst, S Bornemann (2001) Oxalate decarboxylase requires manganese and dioxygen for activity. J Biol Chem 276: 43627–43634. Tanner A, S Bornemann (2000) Bacillius subtilis YvrK is an acid-induced oxalate decarboxylase. J Bacteriol 182: 5271–5273. Tauber MM, A Cavaco-Paulo, K-H Robra, GM Gübitz (2000) Nitrile hydratase and amidase from Rhodococcus rhodochrous hydrolyze acrylic fibers and granular polyacrylonitrile. Appl Environ Microbiol 66: 1634–1638. Teske A, K-U Hinrichs, V Edgcomb, A de Vera Gomez, D Kysela, SP Sylva, ML Sogin, HW Jannasch (2002) Microbial diversity of hydrothermal sediments in the Guaymas Basin: evidence for anaerobic methanotrophic communities. Appl Environ Microbiol 68: 1994–2007. Textor S, VF Wendich, AA De Graaf, U Müller, MI Linder, D Linder, W Buckel (1997) propionate oxidation in Escherichia coli: evidence for operation of a methylcitrate cycle in bacteria. Arch Microbiol 168: 428–436. Thauer RK, D Möller-Zinjhan, AM Spormann (1989) Biochemistry of acetate catabolism in anaerobic chemotrophic bacteria. Annu Rev Microbiol 43: 43–67. Tokimatsu T, Y Nagai, T Hattori, M Shimada (1998) Purification and characteristics of a novel cytochrome c dependent glyoxylate dehydrogenase from a wood-rotting fungus Tyromyces palustris. FEBS Lett 437: 117–121. Toms AV, C Kinsland, DE McCloskey, AE Pegg, SE Ealick (2004) Evolutionary links as revealed by the structure of Thermotoga maritima S-adenosylmethionine decarboxylase. J Biol Chem 279: 33837–33846. Topp E, L Xun, CS Orser (1992) Biodegradation of the herbicide bromoxynil (3,5-dibromo-4-hydroxybenzonitrile) by purified pentachlorophenol hydroxylase and whole cells of Flavobacterium sp. strain ATCC 39723 is accompanied by cyanogenesis. Appl Environ Microbiol 58: 502–506. Uetz T, R Schneider, M Snozzi, T Egli (1992) Purification and characterization of a two-component monooxygenase that hydroxylates nitrilotriacetate from “Chelatobacter” strain ATCC 29600. J Bacteriol 174: 1179–1188. Urakami T, H Araki, H Oyanagi, K-I Suzuki, K Komagata (1990) Paracoccus aminophilus sp. nov. and Paracoccus aminovorans sp. nov., which utilize N,N′-dimethylformamide. Int J Syst Bacteriol 40: 287–291. Urakami T, H Araki, H Oyanagi, K-I Suzuki, K Komagata (1992) Transfer of Pseudomonas aminovorans (den Dooren de Jong 1926) to Aminobacter gen. nov. as Aminobacter aminovorans comb. nov. and description of Aminobacter aganoensis sp. nov. and Aminobacter niigataensis sp. nov. Int J Syst Bacteriol 42: 84–92. van Beilin J, EG Funhoff, A van Loon, A Just, L Kaysser, M Bouza, R Holtackers, M Röthlisberger, Z Li, B Witholt (2006) Cytochrome P450 alkane hydroxylases of the CYP153 family are common in alkane-degrading eubacteria lacking integral membrane alkane hydroxylases. Appl Environ Microbiol 72: 59–65. Van den Tweel WJJ, JAM de Bont (1985) Metabolism of 3-butyl-1-ol by Pseudomonas BB1. J Gen Microbiol 131: 3155–3162. Van den Wijngaard AJ, DB Janssen, B Withold (1989) Degradation of epichlorohydrin and halohydrins by bacterial cultures isolated from freshwater sediment. J Gen Microbiol 135: 2199–2208. Van der Werf MJ, HJ Swarts, JAM de Bont (1999) Rhodococcus erythopolis DCL14 contains a novel degradation pathway for limonene. Appl Environ Microbiol 65: 2092–2102. Van der Werf MJ, KM Overkamp, JAM de Bont (1998) Limonene-1,2-epoxide hydrolase from Rhodococcus erythropolis DCL14 belongs to a novel class of epoxide hydrolases. J Bacteriol 180: 5052–5057. Van Ginkel CG, JAM de Bont (1986) Isolation and characterization of alkene-utilizing Xanthobacter spp. Arch Microbiol 145: 403–407.

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van Hylckama Vlieg JET, H Leemhuis, JHL Spelberg, DB Janssen (2000) Characterization of the gene cluster involved in isoprene metabolism in Rhodococcus sp. strain AD45. J Bacteriol 182: 1956–1963. van Hylckama Vlieg JET, J Kingma, AJ van den Wijngaard, DB Janssen (1998) A gluathione S-transferase with activity towards cis-1,2-dichloroepoxyethane is involved in isoprene utilization by Rhodococcus strain AD 45. Appl Environ Microbiol 64: 2800–2805. van Hylckama Vlieg JET, J Kingma, W Kruizinga, DB Janssen (1999) Purification of a glutathione S-transferase and a glutathione conjugate-specific dehydrogenase is involved in isoprene metabolism by Rhodococcus sp. strain AD 45. J Bacteriol 181: 2094–2101. Walsh C, RA Pascal, M Johnston, R Raines, D Dikshit, A Krantz, M Honma (1981) Mechanistic studies on the pyridoxal phosphate enzyme 1-aminocyclopropane-1-carboxylate deaminase from Pseudomonas sp. Biochemistry 20: 7509–7519. Wang C-S, DA Kunz, BJ Venables (1996) Incorporation of molecular oxygen and water during enzymatic oxidation of cyanide by Pseudomonas fluorescens NCIMB 11764. Appl Environ Microbiol 62: 2195–2197. Warhurst AM, KF Clarke, RA Hill, RA Holt, CA Fewson (1994) Metabolism of styrene by Rhodococcus rhodochrous NCIMB 13259. Appl Environ Microbiol 60: 1137–1145. Wäsch SI de A, JR van der Ploeg, T Maire, A Lebreton, A Kiener, T Leisinger (2002). Transformation of isopropylamine to L-alaninol by Pseudomonas sp. strain KIE171 involves N-glutamylated intermediates. Appl Environ Microbiol 68: 2368–2375. Wayne LG et al. (1991) Fourth report of the cooperative, open-ended study of slowly growing mycobacteria by the international working group on mycobacterial taxonomy. Int J Syst Bacteriol 41: 463–472. Wegener WS, HC Reeves, R Rabin, SJ Ajl (1968) Alternate pathways of metabolism of short-chain fatty acids. Bacteriol Rev 32: 1–26. Weijers CAGM, A de Haan, JAM de Bont (1988) Chiral resolution of 2,3-epoxyalkanes by Xanthobacter Py2. Appl Microbiol Biotechnol 27: 337–340. Widdel F (1986) Growth of methanogenic bacteria in pure culture with 2-propanol and other alcohols as hydrogen donors. Appl Environ Microbiol 51: 1056–1062. Widdel F (1987) New types of acetate-oxidizing, sulfate-reducing Desulfobacter species, D. hydrogenophilus sp. nov., D. latus sp. nov., and D. curvatus sp. nov. Arch Microbiol 148: 286–291. Widdel F, N Pfennig (1981) Sporulation and further nutritional characteristics of Desulfotomaculum acetoxidans (emend). Arch Microbiol 112: 119–122. Widdel F, N Pfennig (1982) Studies on dissimilatory sulfate-reducing bacteria that decompose fatty acids. II. Incomplete oxidation of propionate by Desulfobulbus propionicus gen. nov., sp. nov. Arch Microbiol 131: 360–365. Widdel F, PE Rouvière, RS Wolfe (1988) Classification of secondary alcohol-utilizing methanogens including a new thermophilic isolate. Arch Microbiol 150: 477–481. Wiegant WM, JAM de Bont (1980) A new route for ethylene glycol metabolism in Mycobacterium E44. J Gen Microbiol 120: 325–331. Wilkes H, R Rabus, T Fischer, A Armstroff, A Behrends, F Widdel (2002) Anaerobic degradation of n-hexane in a denitrifying bacterium: further degradation of the initial intermediate (1-methylpentyl)succinate via a C-skeleton rearrangement. Arch Microbiol 177: 235–243. Witschel M, S Nagel, T Egli (1997) Identification and characterization of the two-enzyme system catalyzing the oxidation of EDTA in the EDTA-degrading bacterial strain DSM 9103. J Bacteriol 179: 6937–6943. Yamada EW, WB Jakoby (1959) Enzymatic utilization of acetylenic compounds. II. Acetylenemonocarboxylic acid hydrase. J Biol Chem 234: 941–945. Yamamoto K, Y Ueno, K Otsubo, K Kawakami, K-I Komatsu (1990) production of S-(+)-ibuprofen from a nitrile compound by Acinetobacter sp. strain AK 226. Appl Environ Microbiol 56: 3125–3129. Zeikus JG (1980) Chemical and fuel production by anaerobic bacteria. Annu Rev Microbiol 34: 423–464. Zhou N-Y, A Jenkins, CKN Chan, KW Chion, DJ Leak (1999) The alkene monooxygenase from Xanthobacter strain Py2 is closely related to aromatic monooxygenases and catalyzes aromatic monooxygenation of benzene, toluene, and phenol. Appl Environ Microbiol 65: 1589–1595.

PART 2 CYCLOALKANES The aerobic degradation of cycloalkanes has been examined in both monocyclic and polycyclic substrates. In all of them, monooxygenation is the first step and this is sometimes accomplished by cytochrome P450 systems. Reviews of the degradation of alicyclic compounds including monoterpenes

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Aliphatic Compounds OH

337 O

O O HO2C

FIGURE 7.32

[CH2]4

CO2H

Biodegradation of cyclohexane.

have been given by one of the pioneers (Trudgill 1978, 1984, 1994) and these should be consulted for further details. Only a bare outline with some newer developments will therefore be given here. Hydroxylation by cycloalkane monooxygenase and dehydrogenation to the cycloalkanone are the first steps in the oxidation of cycloalkanes, and are formally similar to those used for the degradation of linear alkanes. It was some time, however, before pure strains of microorganisms were isolated that could grow with cycloalkanes or their simple derivatives. The degradation of cyclohexane has been examined in detail (Stirling et al. 1977; Trower et al. 1985). There are five steps in its degradation: (1) hydroxylation of the ring by monooxygenation, (2) dehydrogenation to the ketone, (3) insertion of one atom of oxygen into the ring in a reaction formally analogous to the Baeyer– Villiger persulfate oxidation, (4) hydrolysis of the lactone, and (5) dehydrogenation to an α,ωdicarboxylate. The pathway is illustrated for cyclohexane in Figure 7.32 (Stirling et al. 1977), and a comparable one operates also for cyclopentanol (Griffin and Trudgill 1972). The enantiomeric specificity of this oxygen-insertion reaction has been examined in a strain of camphor-degrading Pseudomonas putida (Jones et al. 1993). Both cyclohexanone monooxygenase (Branchaud and Walsh 1985) and cyclopentanone monooxygenase exhibit a substrate versatility (Iwaki et al. 2002) that is reminiscent of that of MMO, and has attracted interest in them as biocatalysts. There has been interest in Baeyer–Villiger monooxygenation of higher cycloalkanones and important issues have emerged: a. Cyclododecanone monooxygenase from Rhodococcus ruber SC1 is different from those already mentioned, and is active towards substrates with more than seven carbon atoms (Kostichka et al. 2001). b. The enzyme derived from Pseudomonas sp. strain HI-70 is able to oxidize a wide range of substrates including C12–C15 ketones, C5 and C6 ketones with methyl substituents, and some bicyclic ketones including decalones (Iwaki et al. 2006). Attention has also been directed to the degradation of cycloalkane carboxylic acids: 1. Cyclopropane carboxylic acid is degraded via 3-hydroxybutyrate by both the bacterium Rhodococcus rhodochrous (Toraya et al. 2004) (Figure 7.33) and by fungi (Schiller and Chung 1970), although the mechanism for ring fission has not been determined. 2. The degradation of 11,12-methyleneoctadecanoate is carried out in Tetrahymena pyriformis by a modified β-oxidation pathway with the formation of propionate and acetate (Figure 7.34a) (Tipton and Al-Shather 1974). 3. Cyclohexylacetate is degraded to cyclohexanone by elimination of the side chain after hydroxylation at the ring junction (Figure 7.34b) (Ougham and Trudgill 1982). An alternative to monooxygenation is realized in the hydration of the substituted cyclohexanone derived from the fission product of 1,2-dihydroxynaphthalene during the degradation of 1,2,3,4tetrahydronaphthalene (tetralin) (Hernáez et al. 2002) (Figure 7.35). A less common pathway for the degradation of cyclohexanecarboxylate has been found in which the ring is dehydrogenated to 4-hydroxybenzoate before fission (Figure 7.36) (Blakley 1974; Taylor and Trudgill 1978). The degradation of polyhydroxylated cyclohexanes such as quinate and shikimate also involves aromatic intermediates (Ingledew et al. 1971), though in these examples a

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338

Environmental Degradation and Transformation of Organic Chemicals OH

O CO2H

CO2H CO2H

CO2H

FIGURE 7.33 (a)

Aerobic degradation of cyclopropanecarboxylate. CH 2

CH 3−(CH2)5−CH

CH 2−CO2H

(b) CH−(CH 2)9−CO2H

CH 2 CH 3−(CH2)5−CH

CH−CH 2−CO2H CH−CO 2H

CH 2 CH 3−(CH2)5−CH

C CH−CO 2H CH 2−CO2H

HO

CH 2 CH 3−(CH2)5−CH

CH 2−CO2H C OH

O

+

CH 3−CO2H

CH 3−(CH2)5−CH2−CH2−CO−CH2−CO2H

FIGURE 7.34

Degradation of (a) 11,12-methyleneoctadecanoate and (b) cyclohexylacetate.

OH

OH OH

OH

O HO2C

CO2H OH

OH CO2H CO2H

HO2C OH

FIGURE 7.35

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O

H O2C

CHO + CH3CO−CO2H

Aerobic degradation of tetralin.

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Aliphatic Compounds

339

CO2H

CO2H

CO2H

CO2H

OH OH

O

FIGURE 7.36

Alternative pathway for the biodegradation of a cyclohexane carboxylate. CO 2 H

CO 2 H

HO

HO

OH OH

FIGURE 7.37

OH

HO

CO 2 H

OH OH

CO 2 H

OH

O OH

HO OH

Biodegradation of quinate.

mechanism for the formation of the aromatic ring by elimination reactions is more readily rationalized (Figure 7.37). The interrelation between the metabolism cyclohexane carboxylates and their benzenoid analogs may be seen in the pathways for the anaerobic degradation of cyclohexane carboxylate by Rhodopseudomonas palustris. This takes place by the action of a ligase (AliA) to form the coenzyme ester, followed by a dehydrogenase (BadJ) to produce cyclohex-1-ene-1-carboxy-CoA, which is then fed into the pathway used for the anaerobic degradation of benzoyl-CoA (Egland and Harwood 1999), which is discussed in Chapter 8, Parts 1 and 3. An unusual reaction was been observed in the reaction of “old yellow enzyme” with α,β-unsaturated ketones. A dismutation took place under aerobic or anaerobic conditions, with the formation from cyclohex-1-keto-2-ene of the corresponding phenol and cyclohexanone, and an analogous reaction from representative cyclodec-3-keto-4-enes—putatively by hydride-ion transfer (Vaz et al. 1995). Reduction of the double bond in α,β-unsaturated ketones has been observed, and the enone reductases from Saccharomyces cerevisiae have been purified and characterized. They are able to carry out reduction of the C=C bonds in aliphatic aldehydes and ketones, and ring double bonds in cyclohexenones (Wanner and Tressel 1998). Reductions of steroid 1,4-diene-3-ones can be mediated by the related old yellow enzyme and pentaerythritol tetranitrate reductase, for example, androsta-∆1,4-3,17-dione to androsta-∆4-3,17-dione (Vaz et al. 1995) and prednisone to pregna-∆4-17α, 20-diol-3,11,20-trione (Barna et al. 2001) respectively. Numerous single-step transformations—generally hydroxylations, oxidations of alcohols to ketones or dehydrogenations—of both terpenes and sterols have been accomplished using microorganisms especially fungi. This interest has been motivated by the great interest of the pharmaceutical industry in the products (Smith et al. 1988), and in fungal metabolism as a model for that of higher organisms (Smith and Rosazza 1983). Reactions catalyzed by cytochrome P450 mediated in both prokaryotes and eukaryotes have been discussed in Chapter 3, Part 1. Anaerobic degradation of cycloalkanes has seldom been reported. The pathway used for the degradation of ethylcyclopentane by a sulfate-reducing enrichment is analogous to the fumarate pathway used for n-alkanes (Part 1 of this chapter) with the formation of 3-ethylcyclopentanecarboxylate followed by ring fission to 3-ethylpentan-1,5-dioate (Rios-Harnandez et al. 2003).

MONOTERPENES Hydroxylation and Baeyer–Villiger reactions carried out by monooxygenation are important in the degradation of a range of terpenoids and steroids. The aerobic degradation of limonene can take place by a number of reactions several of which involve hydroxylation at allylic positions

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Environmental Degradation and Transformation of Organic Chemicals

O

O

HO2C

O

O

O

O

O HO2C OH

O

O

O

CO2H HO2C

FIGURE 7.38

Degradation of camphor by Pseudomonas putida.

O

O

O

O

CO2H

FIGURE 7.39

OH

O

O

O

COCH3

CO2H

HO2C

CO2H

Degradation of camphor by Mycobacterium rhodochrous.

(terpene numbering): C-7 to perillyl alcohol, C-6 to carveol, C-4 to α-terpineol, and C-3 to isopiperitenol (references in van der Werf et al. 1999). Oxidations analogous to Baeyer–Villiger reactions are employed in the degradation of alicyclic compounds containing several rings. The degradation of the monoterpenoid camphor has attracted particular attention since the initial hydroxylation in a plasmid-bearing pseudomonad is carried out by a cytochrome P450 enzyme (designated P450cam) (Schlichting et al. 2000). This enzyme displays a wide versatility, which is noted later in the degradation of fluorinated alkanes, of abietic acid, in the hydroxylation of 1,4-fluorobenzene, and in the fungal metabolism of styrene. In Pseudomonas putida ATCC 17453 carrying the CAM plasmid degradation involves an initial cytochrome P450 hydroxylation at C-5 followed by oxidation, and the introduction of an oxygen atom adjacent to the quaternary methyl group (Figure 7.38) (Ougham et al. 1983). In an organism designated Rhodococcus rhodochrous strain NCIMB 9784, hydroxylation occurs, however, at C-6, followed by ring fission of the 1,3-diketone (Figure 7.39) (Chapman et al. 1966). The hydrolase that carries out fission of the cyclic 1,3-diketone has been characterized, and was comparable to those carried out by the enoyl-CoA hydratase superfamily (Grogan et al. 2001; Whittingham et al. 2003). The enzyme from Rhodococcus erythropolis strain DCL14 is able to catalyze insertions of oxygen into the cyclohexanones produced by epoxidation, hydrolysis, and oxidation from both (+)-(4R)- and (−)-(4S)-limonene (van der Werf et al. 1999). In addition, the

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Aliphatic Compounds

341

O

FIGURE 7.40

CHO

Degradation of α-pinene.

monooxygenase from this strain is able to insert an oxygen atom into 2-ketomonoterpenoids at both the 1,2 and 2,3 positions, and into a 3-ketomonoterpenoid at the 3,4 position (terpenoid numbering) (van der Werf 2000). The biodegradation of cyclic monoterpenes has been investigated under both aerobic and denitrifying conditions (Foss and Harder 1998), and may involve key reactions other than the Baeyer–Villiger-type ring cleavage of ketones (Trudgill 1994). For example, although some strains of Pseudomonas sp. degrade α-pinene by rearrangement to limonene, oxidation, and β-oxidation, in others the initial reaction was formation of the epoxide that produced a carbonium ion at C-1 followed by sequential ring fission of the cyclobutane and cyclohexane rings to produce 2-methyl5-isopropylhexa-2,5-dienal (Best et al. 1987; Griffiths et al. 1987a,b) (Figure 7.40). Bacterial hydroxylation by cytochrome P450 systems is well established in actinomyces (O′Keefe and Harder 1991). Stereospecific hydroxylation of α-ionone—though not β-ionone—has been observed with strains of Streptomyces sp. (Lutz-Wahl et al. 1998), and the racemic substrate is hydroxylated to the (3R,6R)and (3R,6S)-hydroxy-α-ionones.

STEROIDS Hydroxylation Microorganisms, particularly fungi, are able to hydroxylate sterols at the 11α-, 12β-, 15-, 17, 19, and 21 positions (references in Hudlický 1990), although the product even from the same substrate may depend on the organism: a. Calonectria decora, Rhizopus nigricans, and Aspergillus ochraceus produce the 12β,15α-; 11α,16β-; and 6β,11α-diols from 3-keto-5α-androstane (Bird et al. 1980). b. The biotransformation of pregna-4,17(20)-cis-diene-3,16-dione by Aspergillus niger produced metabolites with hydroxyl groups at the 7β, or 7β and 15β positions, whereas Cephalosporium aphidicola produces metabolites with hydroxyl groups at the 11α, or 11α and 15β positions (Atta-ur-Rahman et al. 1998). An interesting development has been the need for unusual products that may be sterol metabolites. Studies have been directed to the biosynthesis, for example, of the otherwise rare and inaccessible derivatives of progesterone hydroxylated at the 6, 9, 14, or 15 positions that could be accomplished by incubating progesterone in a complex medium with the fungus Apiocrea chrysosperma (Smith et al. 1988). The hydroxylation of progesterone and closely related compounds at the 15β position has been observed in cell extracts of Bacillus megaterium (Berg et al. 1976). Other Aerobic Transformations and Degradation Attention has been directed to the reactions catalyzed by cytochrome P450s that bring about important reactions leading to the loss of angular methyl groups at C-10 (with concomitant aromatization of ring A) and C-14, and the C-17 –COCH3 side chain. These reactions are discussed further in Chapter 3, Part 1. The transformations of steroids and their precursor lanosterol have been extensively

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342

Environmental Degradation and Transformation of Organic Chemicals CO2H HO (a) O

HO

O

(b) HO

O

HO (c)

A O

A

A O

HO

A

A

A

A

O

HO

O

HO OH

C B

C

C B

OH

B

C B

FIGURE 7.41 Demethylation of angular methyl groups: (a) in steroids and related compounds, (b) in lanosterol derivative and (c) summary of reactions for the formation of CH2O.

examined as sources of estrone and related compounds. In all of them, the essential reactions involve loss of the angular methyl groups and cytochrome P450–mediated reactions have already been noted in Chapter 3, Part 1. In Nocardia sp., demethylation of steroids is accomplished by sequential formation of the 1,4-diene-3-ones, hydroxylation at C-10, and a dienone-phenol transformation (Figure 7.41a) (Sih et al. 1968). In a lanosterol derivative, Mycobacterium sp. carries out analogous reactions for the loss of the C-29, C-30, and C-28 methyl groups by hydroxylation (Figure 7.41b) (Wang et al. 1995). Plausible reactions for the formation of CH2O from the quaternary methyl groups are outlined in Figure 7.41c), although formate is produced in cytochrome P450–mediated reactions. The degradation of sterols and the related bile acids oxygenated at C-3 is also initiated by formation of the 1,4-diene-3-ones. In these, the critical reaction is, however, hydroxylation at C-9 that results in fission of the B-ring with formation of the 9,10-seco compound under the driving force of aromatization of the A-ring by a dienone-phenol reaction (Leppik 1989). The pathway is exemplified in the degradation of testosterone by a strain of Comamonas testosteroni (Horinouchi et al. 2003a, 2005) (Figure 7.42), and is presumptively used for the degradation of androsterone after initial dehydrogenation (Horinouchi et al. 2003b).

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Aliphatic Compounds

343 O

O

O

O O

O

H

O

O

O

O

O

O

OH

O

O

O

HO 2 C OH

OH

OH

FIGURE 7.42

HO

H

Aerobic degradation of testosterone.

OH

O

O

OH

H

O H

FIGURE 7.43

CO 2 H

O

O

HO

OH

O

H

Anaerobic biotransformation of 3α,7α-cholic acid.

Anaerobic Transformation This has been examined quite extensively in the context of the intestinal metabolism of bile acids and a number of reactions that are otherwise quite unusual have been observed, most frequently in organisms belonging to the genera Eubacterium or Clostridium. Illustrative examples include the following: a. Reduction of the ∆4,5 bond with production of 5β-reduced compounds (Mott et al. 1980). b. Reductive dehydroxylation of 7α-hydroxy bile acids (Masuda et al. 1984) that has been extensively studied, and both the genetics and the complex of pathways have been elucidated for cholic acid (Wells and Hylemon 2000). These reactions involve a complex sequence of reactions: (i) dehydrogenation to the 3-keto-4,5-ene, (ii) dehydration to the 3-keto-4,6-diene, and (iii) successive reduction to the 3-keto and the 3-hydroxy compounds, which are illustrated in Figure 7.43 (Wells and Hylemon 2000). In addition, dehydroxylation of 16α- and 21-hydroxy corticosteroids has been observed with fecal bacteria (Bokkenheuser et al. 1980).

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Environmental Degradation and Transformation of Organic Chemicals

(a)

A

B

B

A

A

B O

OH CO 2 H

CO 2 H

OH

CO 2 H

HO O A

B

A

CO 2 H

B

O

O

CO 2 H

CO 2 H CO 2 H

CO 2 H

OH HO

(b)

A

A

B

B

O

A

B O

CO 2 H

A HO

FIGURE 7.44

CO 2 H

Pathways for aerobic degradation of dehydroabietic acid.

DITERPENOIDS The degradation of the diterpenoid derivative dehydroabietic acid has been examined in several different strains of bacteria. In Pseudomonas abietaniphila strain BKME-9, it is degraded by initial cytochrome P450 hydroxylation at C-7 (Smith et al. 2004), oxidation to the 7-keto compound followed by dioxygenation to the 7-keto-11,12-diol that undergoes meta fission (Martin and Mohn 1999, 2000). Further degradation then takes place by dioxygenation of ring C and ring fission (Figure 7.44a) (Biellmann et al. 1973b) or by an alternative (Figure 7.44b) (Biellmann et al. 1973a).

POLYCYCLIC SUBSTRATES The following illustrate the variety of hydroxylations of structurally diverse polycyclic substrates: 1. P450cam is able to hydroxylate the –CH3 group of the quaternary methyl group of 5,5difluorocamphor (Figure 7.45a) to the 9-hydroxymethyl compound (Eble and Dawson 1984), and both adamantane and adamantan-4-one at the -CH quaternary carbon atom (Figure 7.45b) (White et al. 1984). 2. Patchoulol is transformed by Botrytis cinerea to a number of products principally to those involving hydroxylation at the C-5 and C-7 quaternary atoms (Aleu et al. 1999) (Figure 7.45c).

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Aliphatic Compounds

345 OH

O

O

(a) F

F

F

F OH

(b)

14

HO

1

13

4

5

12

(1 )

FIGURE 7.45

OH

HO

6

10 9

HO

3 7

8

(c)

15 2

11

OH (2 )

(3)

Hydroxylation of (a) 5,5-difluorocamphor, (b) adamantane, and (c) patchoulol.

HO

O

N

N CH 2

FIGURE 7.46

O CH 2

Hydroxylation of azabrendane derivative.

3. Cells of Acinetobacter sp. NCIB 9871 grown with cyclohexanol carried out enantiomerically specific degradation of 5-bromo-7-fluoronorbornanone and production of a lactone with >95% enantiomeric excess (Levitt et al. 1990). 4. Bauveria sulfurescens stereospecifically hydroxylated an azabrendane at the quaternary carbon atom (Figure 7.46) (Archelas et al. 1988). 5. Penicillium lilacinum transformed testosterone successively to androst-4-ene-3,17-dione and testololactone (Prairie and Talalay 1963): once again the oxygen atom is introduced into ring D at the quaternary position between C13 and C17. 6. An illustration of the plethora of reactions that may occur is afforded by the transformation of caryophyllene oxide by Botrytis cinerea. Although most of the reactions were hydroxylations or epoxidations, two involved transannular reactions: (a) between the C4-epoxide oxygen and C7 and (b) between the C4 -epoxide and C13 with formation of a caryolane (Figure 7.47) (Duran et al. 1999). 7. The degradation of atropine has been examined in Pseudomonas sp. strain AT3 and produces tropine as the initial metabolite. The degradation of this proceeds by oxidative

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O O

OH +

HO

FIGURE 7.47

Transformation of caryophyllene oxide by Botrytis cinerea.

CH 3

H

N

N

O

O

O

O

O

CO 2 H CH 3

OH

FIGURE 7.48

OH

OH

O

O

Degradation of tropine.

loss of the N-methyl group and elimination of ammonia to form 6-hydroxy-cyclohepta-1,4dione, followed by fission of the 1,3-diketone to produce 4,6-diketoheptanoate (Figure 7.48) (Bartholomew et al. 1996). Although the enzymology was unresolved, loss of ammonia presumably occurs either by successive hydroxylation at the tertiary carbon atoms adjacent to the –NH group or by successive dehydrogenations. Hydroxylation of other natural products such as a. A synthetic diterpenoid butenolide by Cunninghamella elegans at 5α, and both 7α and 7β positions as well as on the isopropyl side chain (Milanova et al. 1994). b. Sclareol by Cunninghamella elegans at the 2α, 3β, 18, and 19 positions (Abraham 1994). c. Stemodin by Cephalosporium aphidicola at the 7α 7β, 8β, 18, and 19 positions (Hanson et al. 1994). The spectrum of substrates has been extended to xenobiotics, and a single example may suffice. The fungus Beauveria bassiana metabolized the hydroxycoumarin rodenticide warfarin to the hydroxylated and oxidized metabolites, which have already been established in mammalian systems or mediated by fungal cytochrome P450 monooxygenase systems. In addition, novel products were produced from the reduction of the keto group in the 4-hydroxyl group (Griffiths et al. 1992).

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Iwaki H, S Wang, S Grosse, H Bergeron, A Nagahashi, J Lertvorachon, J Yang, Y Konishi, Y Hasegawa, PCK Lau (2006) Pseudomonad cyclopentadecanone monooxygenase displaying an uncommon spectrum of Baeyer-Villiger oxidations of cyclic ketones. Appl Environ Microbiol 72: 2707–2720. Iwaki H, Y Hosegawa, S Wang, MM Kayser, PCK Lau (2002) Cloning and characterization of a gene cluster involved in cyclopentanol metabolism in Comamonas sp. strain NCIMB 9872 and biotransformations effected by Escherichia coli-expressed cyclopentanone 1,2-monooxygenase. Appl Environ Microbiol 68: 5671–5684. Jones KH, RT Smith, PW Trudgill (1993) Diketocamphane enantiomer-specific ‘Bayer-Villiger’ monooxygenases from camphor-grown Pseudomonas putida ATCC 17453. J Gen Microbiol 139: 797–805. Kostichka K, SM Thomas, KJ Gibson, V Nagarajan, Q Cheng (2001) Cloning and characterization of a gene cluster for cyclododecanone oxidation in Rhodococcus ruber SC1. J Bacteriol 183: 6478–6486. Leppik RA (1989) Steroid catechol degradation: disecoandrostane intermediates accumulated by Pseudomonas transposon mutant strains. J Gen Microbiol 135: 1979–1988. Levitt MS, RF Newton, SM Roberts, AJ Willetts (1990) Preparation of optically active 6′-fluorocarbocyclic nucleosides utilising an enantiospecific enzyme-catalysed Baeyer-Villiger type oxidation. J Chem Soc Chem Commun 619–620. Lutz-Wahl S, P Fischer, C Schmidt-Dannert, W Wohlleben, B Hauer, RD Schmid (1998) Stereo- and regioselective hydroxylation of α-ionone by Streptomyces strains. Appl Environ Microbiol 64: 3878–3881. Martin VJJ, WW Mohn (1999) A novel aromatic ring-hydroxylating dioxygenase from the diterpenoiddegrading Pseudomonas abietaniphila BKME-9. J Bacteriol 181: 2675–2682. Martin VJJ, WW Mohn (2000) Genetic investigation of the catabolic pathway for degradation of abietane diterpenoids by Pseudomonas abietaniphila BKME-9. J Bacteriol 182: 3784–3793. Masuda N, H Oda, S Hirano, M Masuda, H Tanaka (1984) 7α-dehydroxylation of bile acids by resting cells of a Eubacterium lentum-like intestinal anaerobe, strain c-25. Appl Environ Microbiol 47: 735–739. Milanova R, M Moore, Y Hirai (1994) Hydroxylation of synthetic abietane diterpenes by Aspergillus and Cunninghamella species: novel route to the family of diterpenes isolated from Tripterygium wilfordii. J Nat Prod 57: 882–889. Mott GE, AW Brinkley, CL Mersinger (1980) Biochemical characterization of cholesterol-reducing Eubacterium. Appl Environ Microbiol 40: 1017–1022. O′Keefe DP, PA Harder (1991) Occurrence and biological function of cytochrome P-450 monoxygenase in the actinomycetes. Mol Microbiol 5: 2099–2105. Ougham HJ, DG Taylor, PW Trudgill (1983) Camphor revisited: involvement of a unique monooxygenase in metabolism of 2-oxo-∆3-4,5,5-trimethylcyclopentenylacetic acid by Pseudomonas putida. J Bacteriol 15: 140–152. Ougham HJ, PW Trudgill (1982) Metabolism of cyclohexaneacetic acid and cyclohexanebutyric acid by Arthrobacter sp. strain CA1. J Bacteriol 150: 1172–1182. Prairie RL, P Talalay (1963) Enzymatic formation of testololactone. Biochemistry 2: 203–208. Rios-Harnandez LA, LM Gieg, JM Suflita (2003) Biodegradation of an alicyclic hydrocarbon by a sulfatereducing enrichment from a gas condensate-contaminated aquifer. Appl Environ Microbiol 69: 434–443. Schiller JG, AE Chung (1970) Mechanism of cyclopropane ring cleavage in cyclopropanecarboxylic acid. J Biol Chem 245: 6553–6557. Schlichting I, J Berendzen, K Chu, AM Stock, SA Maves, DE Benson, RM Sweet, D Ringe, GA Petsko, SG Sligar (2000) The catalytic pathway of cytochrome P450cam at atomic resolution. Science 287: 1615–1622. Sih CJ, KC Wang, HH Tai (1968) Mechanisms of steroid oxidation by microorganisms XIII. C22 acid intermediates in the degradation of the cholesterol side chain. Biochemistry 7: 796–807. Smith DJ, VJJ Martin, WW Mohn (2004) A cytochrome P450 involved in the metabolism of abietane diterpenoids by Pseudomonas abietaniphila BKME-9. J Bacteriol 186: 3631–3639. Smith KE, S Latif, DN Kirk, KA White (1988) Microbial transformations of steroids—I. Rare transformations of progesterone by Apiocrea chrysosperma. J Steroid Biochem 31: 83–89. Smith RV, JP Rosazza (1983) Microbial models of mammalian metabolism. J Nat Prod 46: 79–91. Stirling LA, RJ Watkinson, IJ Higgins (1977) Microbial metabolism of alicyclic hydrocarbons: isolation and properties of a cyclohexane-degrading bacterium. J Gen Microbiol 99: 119–125. Taylor DG, PW Trudgill (1978) Metabolism of cyclohexane carboxylic acid by Alcaligenes strain W1. J Bacteriol 134: 401–411. Tipton CL, NM Al-Shathir (1974) The metabolism of cyclopropane fatty acids by Tetrahymena pyriformis. J Biol Chem 249: 886–889.

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PART 3 ALKANES, CYCLOALKANES AND RELATED COMPOUNDS WITH CHLORINE, BROMINE, OR IODINE SUBSTITUENTS Particularly the chlorinated compounds have enjoyed range of applications: vinyl chloride (chloroethene) as monomer for the production of PVC, tetra- and trichloroethenes as solvents for degreasing, and the insecticides 1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane (DDT) and isomers of hexachlorocyclohexane (HCH) (benzene hexachloride). The biodegradation of fluorinated aliphatic compounds is generally different from the outlines that have emerged from investigations on their chlorinated, brominated, and even iodinated analogues. They are therefore treated separately in Part 4 of this chapter.

CHLORINATED, BROMINATED, AND IODINATED ALKANES, ALKENES, AND ALKANOATES A range of mechanisms has been found for the biodegradation of halogenated alkanes, alkenes, and alkanoates: 1. Elimination including dehydrohalogenation 2. A corrinoid pathway for C1 halides, which operates under both aerobic and anaerobic conditions 3. Nucleophilic displacement including hydroxylation, hydrolysis, and glutathione-mediated reactions 4. Reduction in which a halogen atom is replaced by hydrogen including dehalorespiration 5. Cytochrome P450–mediated reactions.

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Elimination Reactions In contrast to chemical reactions in which elimination and nucleophilic displacement are alternatives and may occur simultaneously, microbial elimination is less common. This term is also used for the reaction in which, for example, 1,2-dihaloethanes are transformed to ethene, in contrast to dehydrohalogenation in which a haloethene is produced or reductive hydrogenolysis to a haloethane. Degradation involving elimination is found in several degradations: a. Several steps in the degradation of DDT by the facultatively anaerobic bacterium Aerobacter aerogenes involve elimination (Figure 7.49) (Wedemeyer 1967), while a range of products from transformation of the trichloromethyl groups have been isolated (Schwartzbauer et al. 2003). The recovery of the elimination product 1,1-dichloro-2,2-bis(p-chlorophenyl)ethene (DDE) from environmental samples long after restriction on the use of DDT suggests a high degree of persistence. DDE can, however, be degraded by ring dioxygenation and extradiol ring fission to 4-chlorobenzoate in cells of Pseudomonas acidovorans M3GY grown with biphenyl (Hay and Focht 1998) and Terrabacter sp. strain DDE-1 induced with biphenyl (Aislabie et al. 1999). An alternative pathway for the transformation of DDT involves hydroxylation of the ring and displacement of the aromatic ring chlorine atom by hydroxyl (Figure 7.50) (Massé et al. 1989). Bis(p-chlorophenyl)acetic acid (DDA) is a polar metabolite that is apparently persistent in the environment (Heberer and Dünnbier 1999). In higher organisms, DDE can be metabolized by formation of sulfones (Letcher et al. 1998). b. Growth of a range of bacteria including Escherichia coli, Bacillus subtilis, and Streptococcus pyogenes is inhibited by 3-chloro-d-alanine and extracts of the first two of these inactivate alanine racemase (Manning et al. 1974). The enzyme that brings about degradation by elimination to pyruvate, ammonia, and chloride has been purified from Pseudomonas putida strain CR 1-1 and contains 2 mol pyridoxal 5′phosphate per mole of enzyme. The enzyme is induced by 3-chloro-d-alanine and also catalyzes the analogous degradation of d-cysteine to pyruvate, sulfide, and ammonia, although this activity could not be detected in medium supplemented with d-cysteine (Nagasawa et al. 1982). There are important differences in the effects of d- and l-3-chloroalanine on membrane transport of proline that are reflections of their metabolism. Whereas in membrane vesicles from E. coli B, the l-enantiomer can undergo elimination to pyruvate catalyzed by pyridoxal phosphate, the d-enantiomer brings about irreversible inactivation of membrane transport by dehydrogenation to chloropyruvate (Kaczorowski et al. 1975). Aerobic Conditions A great deal of effort has been expended in elucidating the details of the degradation of HCH isomers that differ both in their toxicity and their biodegradability. This has been examined mostly in strains originally classified as Sphingomonas paucimobilis, and to a lesser extent in Rhodobacter lindaniclasticus (Nalin et al. 1999), though this strain apparently no longer exists. Although many of the reactions are biotransformations in which the ring is retained, evidence for fission of the ring has been obtained for Sphingobium japonicum strain UT26 that is noted below. 1. Bacteria that degrade or transform HCH isomers have been described. They vary in selectivity towards the isomers and the β-isomer that has only equatorial substituents is generally the most recalcitrant. Three strains, which have been most studied belong to the genus Sphingobium, Sph. indicum B90A, Sph. japonicum UT26, and Sph. francese sp+ (Pal et al. 2005). All have the genes for linA, linD, linC, and linDER, which are discussed later, but they differ in the ease with which they transform the HCH isomers (Table 7.1) (Sharma et al. 2006).

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FIGURE 7.49

Cl

Cl

Cl

Cl

Biotransformations of DDT.

DDMS

CH 2 Cl

CH

DDT

CCl3

CH

Cl

Cl

Cl

Cl

DDA

CO 2 H

CH

DDNU

CH 2

C

DDD

CHCl2

CH

CCl2

C

DDE

Cl

Cl

Cl

Cl

Cl

CO

Cl

Cl

DDMU

CHCl

C

Cl

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Environmental Degradation and Transformation of Organic Chemicals CCl3 CH

Cl

CCl3 Cl

Cl

OH

CH

OH

OH

FIGURE 7.50 Alternative pathway for biodegradation of DDT.

TABLE 7.1 Relative Transformation of Hexachlorocyclohexane Isomers Hexachlorocyclohexane Isomer Sphingobium Strain α γ B90A +++ +++ sp+ ++ ++ UT26 ++ ++ A strain of Pandoraea sp. was also able to degrade the γ-isomer (Okeke et al. 2002).

β +++ ++ +

δ +++ ++ ++

2. There are seven isomers of HCH that differ in their conformations: α-isomer, aaaaee; β-isomer, eeeeee; γ-isomer, aaaeee; δ-isomer, aeeeee; ε-isomer, aeeaee; η-isomer, aaeaee; and φ-isomer, aeaeee (Willett et al. 1998). Their degradation is initiated by dehydrochlorination that is dependent on the stereochemistry of the isomer. For example, the β-isomer that lacks 1,2-diaxial hydrogen and chlorine groups is not a substrate for the LinA dehydrohalogenase (Nagata et al. 1993) and as a consequence this isomer is less readily degraded. Strain Sphingobium japonicum UT26 is, however, capable of converting the persistent β-isomer only into pentachlorocyclohexanol by the activity of the dehalogenase lin B (Nagata et al. 2005). 3. In strains of S. paucimobilis, most of the genes (linA-E) encoding the degradative enzymes are associated with an insertion sequence (IS6100) that mediated horizontal gene transfer (Dogra et al. 2004). In another strain that is able to degrade γ-HCH, the enzymes involved in dehydrochlorination were apparently extracellular (Thomas et al. 1996). 4. a. The first two steps in the biotransformation of HCH involve eliminations (dehydrochlorinations), for example, in the degradation of γ-hexachloro(aaaeee)cyclohexane from which pentachlorobenzene (Tu 1976) or γ-2,3,4,6-tetrachlorocyclohex-1-ene are formed (Jagnow et al. 1977) (Figure 7.51). The formation of 2,5-dichlorophenol and 2,4,5-trichlorophenol during the aerobic degradation of γ-hexachloro(aaaeee)cyclohexane by Sphingomonas (Pseudomonas) paucimobilis putatively involves comparable elimination reactions (Sendoo and Wada 1989). The transformation by Sphingobium japonicum (Pseudomonas paucimobilis) strain UT26 produced 1,2,4-trichlorobenzene as a terminal metabolite by sequential diaxial dehydrochlorinations (Nagasawa et al. 1993a,b). Further details have been given of the stereochemistry of the initial reactions that involve 1,2-diaxial dehydrochlorination to pentachlorocyclohexane followed by sequential 1,4-anti dehydrochlorinations to 1,2,4,5-tetrachlorocyclohex-2,5-diene and 1,2,4-trichlorobenzene (Figure 7.52) (Trantírek et al. 2001). b. In the degradation by Sphingobium japonicum strain UT26, the genes leading to 2,5-dichlorohydroquinone have been designated linA, linB, and linC. They encode enzymes that carry out two successive dehydrochlorinations (LinA), two successive dehalogenations (hydrolytic dechlorinations) (LinB), and a dehydrogenase (LinC) (Nagata et al. 1999). Further steps involving linD, linE, and linF in this strain have been described. They comprise dioxygenation of 2-chlorohydroquinone, followed by

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Aliphatic Compounds

353 Cl

Cl

Cl

Cl

Cl

Cl Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl Cl

FIGURE 7.51

Pathway for biotransformation of γ-hexachlorocyclohexane.

H

H

Cl

H Cl

Cl Cl

FIGURE 7.52

Cl

Cl

Cl

H

Cl

Cl

Cl

H

H

H

H

Cl H

H Cl

H Cl

H

Cl

Cl

Cl

Stereochemistry of first steps in the biotransformation of γ-hexachlorocyclohexane.

hydrolysis of the acyl chloride to 3-hydroxymuconate that is mediated by an unusual metacleavage enzyme encoded by linE (Figure 7.53) (Miyauchi et al. 1999; Endo et al. 2005). Since the elimination reactions that are involved are not themselves dependent on the presence of oxygen, they may occur under anaerobic conditions. 5. Whereas strain UT26 is capable of converting the β-isomer of HCH only into pentachlorocyclohexanol by the activity of the dehalogenase lin B (Nagata et al. 2005), Sphingobium indicum (Sphingomonas paucimobilis) strain B90A is able to degrade all four isomers (α-, β-, γ-, and δ-) of HCH mediated by linA1, linA2, linB, linC, linD, and linE (Suar et al. 2004). 6. Strain B90A contains two copies of linA that have important consequences. The corresponding dehydrochlorinases LinA1 and LinA2 are specific for the (+) and (−) enantiomers of α-HCH (Suar et al. 2005), and the presence of both makes possible the degradation of racemic HCH. In strains with only a single copy of the enzyme, however, degradation of only one of the enantiomers will take place with consequent enrichment of the nondegradable enantiomer. Anaerobic Conditions Elimination is rather exceptional under anaerobic conditions although there are some examples that involve additionally dehydrogenation. 1. Elimination, dehydrogenation, and hydrogenolysis may take place during biotransformation of HCH isomers. a. A strain of Clostridium rectum (Ohisa et al. 1980) converted γ-HCH into 1,2,4-trichlorobenzene, while γ-1,3,4,5,6-pentachlorocyclohexene produced 1,4-dichlorobenzene (Figure 7.54). It was suggested (Ohisa et al. 1982) that this reductive dechlorination and

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Cl LinE

LinA

HO

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

CO 2 H

OH

Cl

Cl

COCl HO

Cl

Cl

CO 2 H

LinB

Cl

CO 2 H

Cl

Cl

LinA

Cl

FIGURE 7.53 Aerobic degradation of γ-hexachlorocyclohexane.

OH

OH

Cl

Cl

Cl

Cl

Cl

Cl

LinF

LinB

O

Cl

Cl

Cl

Cl

CO 2 H

CO 2 H

OH

OH

Cl

LinC Cl OH

OH Cl LinD

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Aliphatic Compounds

355

Cl

Cl

Cl

Cl

Cl

Cl

Cl Cl

Cl

Cl ;

Cl

Cl

Cl

Cl Cl

Cl

FIGURE 7.54 Anaerobic biotransformation of γ-hexachlorocyclohexene and γ-1,3,4,5,6-pentachlorocyclohexene. Br CH2 CH2 Br

FIGURE 7.55

CH2

CH2

;

Br CH

CH Br

CH

CH

Metabolism of 1,2-dibromoethane.

elimination was coupled to the synthesis of ATP, and this possibility has been amply confirmed in a number of anaerobic dehalogenations of polychlorinated ethenes and some chlorophenols. b. During anaerobic incubation of [36Cl]-γ-hexachlorocyclohexane, dechlorination with tetrachlorocyclohexene as an intermediate was demonstrated in Cl. butyricum, Cl. pasteurianum, and Citrobacter freundii (Jagnow et al. 1977). c. A methanogenic enrichment culture dehydrochlorinated β-hexachlorocyclohexane to δ-tetrachlorocyclohexene by elimination and hydrogenolysis, with the formation of chlorobenzene and benzene as the stable end product. The α-isomer was dechlorinated at a comparable rate, and the γ- and δ-isomers more slowly (Middeldorp et al. 1996). 2. Analogous reactions have been shown to occur with aliphatic halides: a. Pure cultures of strictly anaerobic methanogenic bacteria transformed 1,2-dibromoethane to ethene and 1,2-dibromoethene to ethyne (Figure 7.55) (Belay and Daniels 1987). b. A nonmethanogenic culture produced propene from 1,2-dichloropropane plausibly via the reductive formation of monochloropropanes followed by elimination (Löffler et al. 1997). c. The dehalorespiring Desulfitobacterium dichloroeliminans strain DCA1 carried out exclusive antielimination of vicinal dichloroethane, and all the vicinal dichlobutanes (de Wildeman et al. 2003). d. The reductive dehalogenase from Desulfitobacterium sp. strain Y51 carried out not only the dehalogenation of tetrachloroethene to trichloroethene and cis-1,2-dichloethene, but also the elimination of a number of polyhalogenated ethanes, for example, hexachloroethane to cis-1,2-dichloroethene (Suyama et al. 2002). Corrinoid Pathways The existence of corrinoids in anaerobic bacteria in substantial concentrations is well established, and their metabolic role in acetogenesis and in methanogenesis has been elucidated. Their involvement in degradation pathways of aerobic organisms is more recent, and it has emerged that their role under these different conditions is similar. These issues are explored in the following paragraphs with a view to illustrating the similar metabolic pathways used by both aerobes and anaerobes. Corrinoids are involved in aerobic degradation of methyl chloride by the aerobic Methylobacterium sp. strain CM4, and also in the degradation of other C1 compounds by Methylobacterium extorquens (Christoserdova et al. 1998). Methyl corrins are key components in transmethylation and examples illustrating the similarity of pathways in aerobic and anaerobic metabolism will be summarized. In the following discussion, THF or tetrahydromethanopterin (Figure 7.56) are implicated in the form of their methyl (CH3), methylene (CH2), methine (CH), and formyl (CHO) derivatives (Figure 7.57). The formation of a CH3–Co bond is central to activity, and generally the 5,6-dimethylbenziminazole is replaced by histidine.

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356

Environmental Degradation and Transformation of Organic Chemicals H

O

H2N

CH 2

N

HN

NHΦ

N H Tetrahydrofolate H4F

H N

3

2

N

N

HN H2N

N

4

CH 3

O

N

CH

NHΦ

N H

Tetrahydromethanopterin H 4 MPT N

5

10

CH 2 9 H 4 F 6

N 1

8

7

N H

FIGURE 7.56 Partial structures of tetrahydrofolate (H4F) and tetrahydromethanopterin (H4MPT). (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, pp. 1–74, Springer, 2002. With permission.)

O N H 2N

H N

CH 2 NHΦ H 4F

H

N H CH 3 N

CH 2 NHΦ CH 3

H4F

CH 2

H4F

N H CH 2 N N

Φ

CH 2

N H Φ

CH

N

N

CH 2 CH

H 4F

N H CHO N

CH 2 NHΦ CHO

H4F

N H

FIGURE 7.57 Dehydrogenation of CH3-tetrahydrofolate to CHO-tetrahydrofolate. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, pp. 1–74, Springer, 2002. With permission.)

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Aliphatic Compounds

357

Strain IMB-1 is able to grow at the expense of methyl bromide (Woodall et al. 2001) and belongs to a group of organisms that can also degrade methyl iodide, but are unable to use formaldehyde or methanol (Schaefer and Oremland 1999). It was postulated that the pathway for chloromethane degradation in this strain was similar to that in Methylobacterium chloromethanicum (McAnulla et al. 2001a). Aerobic Degradation of Methyl Chloride Methylotrophic bacteria have been isolated that are able to use methyl chloride aerobically as the sole source of energy and carbon, and bacteria that can utilize methyl chloride are apparently widely distributed and several belong to strains of Hyphomicrobium (McAnulla et al. 2001b). The substrate is metabolized to formaldehyde and subsequently oxidized either to formate and CO2, or incorporated via the serine pathway. A study using strain CC495 that is similar to the strain IMB-1 already noted revealed the complexity of this reaction (Coulter et al. 1999), while details had emerged from a somewhat earlier study of methyl chloride degradation by the aerobic Methylobacterium sp. strain CM4 (Methylobacterium chloromethanicum). Cobalamin was necessary for growth with methyl chloride, though not for growth with methylamine, and the use of mutants containing a miniTn5 insertion and enzyme assays revealed that the mechanism involved initial methyl transfer to a Co(I) corrinoid followed by oxidation via THFs to formytetrahydrofolate and thence to formate with production of ATP (Figure 7.58) (Vannelli et al. 1999). Anaerobic Degradation of Methyl Chloride The anaerobic methylotrophic homoacetogen Acetobacterium dehalogenans is able to grow with methyl chloride and CO2, and uses a pathway comparable to that noted above for the aerobic

CH 3 Cl CmuA III

CH 3

Co

CH 3

H 4F

CH 2

H 4F

E

CmuB

MetF

FolD CH

H 4F

FolD CHO

H 4F

PurU 2H

Methylobacterium chloromethanicum

FIGURE 7.58 Degradation of methyl chloride by Methylobacterium chloromethanicum. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, pp. 1–74, Springer, 2002. With permission.)

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358

Environmental Degradation and Transformation of Organic Chemicals CH 3 Cl

CH 3

H4F

CH 2

H4F

CH 3

Co I

E

CO 2 CO

CH

H4F

CHO

H4 F

CH 3CO−SCoA

CH 3 CO 2 H

H−CO 2H

FIGURE 7.59

Degradation of methyl chloride by Acetobacterium halogenans.

degradation: dehydrogenation of the methyl group involving THF and a corrinoid coenzyme. Acetate is simultaneously produced from CO2 by the activity of CO dehydrogenase and the methyltetrahydrofolate (Figure 7.59) (Meβmer et al. 1993). Some of the gene products are shared with those involved in metabolism of methyl chloride (Vannelli et al. 1999). The methyl transfer reactions and those involved in the subsequent formation of acetate have been explored for the demethylase of this strain (Kaufmann et al. 1998), and resemble closely those for the aerobic metabolism of methyl chloride by aerobic methylotrophs. Anaerobic Degradation of Polyhalogenated Methanes Acetobacterium dehalogenans is able to degrade dichloromethane and the pathway formally resembles that for the anaerobic degradation of methyl chloride. A strain of Dehalobacterium formicoaceticum is able to use only dichloromethane as a source of carbon and energy forming formate and acetate (Mägli et al. 1998). The pathway involves initial synthesis of methylenetetrahydrofolate of which two-thirds is degraded to formate with generation of ATP, while the other third is dehydrogenated, transmethylated, and after incorporation of CO forms acetate with production of ATP (Figure 7.60). The formation of [13C]formate, [13C]methanol, and [13CH3]CO2H was elegantly confirmed using a cell suspension and [13C]CH2Cl2. It was suggested that a sodium-independent F0F1-type ATP synthase exists in this organism in addition to generation of ATP from formyltetrahydrofolate. A strain of Acetobacterium woodii strain DSM 1930 dehalogenated tetrachloromethane to dichloromethane as the final chlorinated product, while the carbon atom of [14C]tetrachloromethane was recovered as acetate (39%), CO2 (13%), and pyruvate (10%) (Egli et al. 1988). Since the transformation of tetrachloromethane to chloroform and CO2 is a nonenzymatic corrinoid-dependent reaction (Egli et al. 1990; Hasham and Freedman 1999), it seems safe to assume operation of the acetyl-CoA synthase reaction. The synthesis of acetate that also takes place during the degradation of dichloromethane by Dehalobacterium formicoaceticum involves CO2 that originates from the medium (Mägli et al. 1996). Nucleophilic Substitution: Hydrolytic Reactions of Halogenated Alkanes and Alkanoates Alkanes Displacement of halogen by hydroxyl is a widely distributed reaction in the degradation of haloalkanes and haloalkanoates. Although an apparently simple pathway involving two displacement

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Aliphatic Compounds

359 CH2Cl2 + H4F

CH 2

H 4F

CH 2

H 4F

CH

H 4F

CH 3

H 4F

CHO

H 4F

CH 3

Co I

E

CO 2 CO

H−CO 2H CH 3 C O SCoA

CH 3 C O PO 4

CH3CO2H

FIGURE 7.60 Degradation of methylene chloride by Dehalobacterium formicoaceticum. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, pp. 1–74, Springer, 2002. With permission.) Cl−CH 2−CH2−Cl

FIGURE 7.61

Cl−CH 2−CH2OH

Cl−CH 2CHO

Cl−CH 2−CO2H

HOCH 2−CO2H

Biodegradation of 1,2-dichloroethane.

steps is illustrated in Figure 7.61 (Janssen et al. 1985), it should be emphasized that the enzymology of hydrolytic dehalogenation is quite complex. For example, two different dehalogenases are involved in the dechlorination of 1,2-dichloroethane and chloroacetate (van den Wijngaard et al. 1992). A number of distinct dehalogenases exist, and they may differ significantly in their substrate specificity in respect of chain length and the influence of halogen atoms at the ω-position (Scholtz et al. 1988; Sallis et al. 1990). The degradation of 1,2-dichloroethane may also be initiated by monooxygenation (Hage and Hartmans 1999), and the pathways for hydrolysis and monooxygenation converge with the production of glyoxylate. The degradation of 2-chloroethylvinyl ether by Ancylobacter aquaticus is initiated by a dehalogenase, although fission of the C–O–C bond is nonenzymatic (van den Wijngaard et al. 1993). Pseudomonas sp. strain ES-2 was able to grow with a range of brominated alkanes that greatly exceeded the range of chlorinated or unsubstituted alkanes, and bromoalkanes with chain lengths of C6 –C16 and C18 could be utilized (Shochat et al. 1993). A range of chlorinated, brominated, and iodinated alkanes C4 –C16 was incubated with resting cells of Rhodococcus rhodochrous NCIMB 13064 (Curragh et al. 1994), and dehalogenation was assessed from the concentration of halide produced. The range of substrates is impressive and the yields were approximately equal for chloride and bromide, and greater for iodide. Significant differences between the degradation of

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Environmental Degradation and Transformation of Organic Chemicals

chlorine- and bromine-substituted alkanes may, however, exist. For example, although Xanthobacter autotrophicus is able to grow with 1,2-dichloroethane and the dehalogenases can debrominate 1,2dibromoethane and bromoacetate, these substrates are unable to support growth of the organism. Several reasons have been suggested including the toxicity of bromoacetaldehyde (van der Ploeg et al. 1995). This is consistent with the observations that in this strain, the initially produced 2chloroethanol is oxidized to the aldehyde by an alkanol dehydrogenase and then to chloroacetate before loss of chloride and mineralization. In contrast, Mycobacterium sp. strain GP1,which belongs to the group of fast-growing mycobacteria, was able to use 1,2-dibromoethane as a source of carbon and energy. Although a hydrolytic haloalkane dehalogenase produced 2-bromoethanol, this was converted into the epoxide that was used for growth by a pathway that was not established. In this way, production of toxic bromoacetaldehyde was circumvented (Poelarends et al. 1999). The metabolic capacity of species of mycobacteria including the human pathogen Mycobacterium tuberculosis strain H37Rv is remarkable (Jesenská 2000). Extracts of M. avium and M. smegmatis were able to dehalogenate a range of halogenated alkanes with chlorine, bromine, and iodine terminal substituents. On the basis of amino acid and DNA sequences, the strain that was used contained three halohydrolases, and the debromination capability of a selected number of other species of mycobacteria is given in Table 7.2. The haloalkane dehalogenase gene from M. avium has been cloned and partly characterized (Jesenská et al. 2002), and it has been shown that there are two dehalogenase genes dmbA and dmbB that are widely distributed in bacteria of the Mycobacterium tuberculosis complex (Jesenská et al. 2005). Some strains of bacteria are able to use α,ω-dichlorinated long-chain alkanes for growth, and the activity of the hydrolase from Rhodococcus erythropolis strain Y2 was high for 1,2-dibromoethane, 1,2-dibrompropane, and the α,ω-dichloroalkanes (Sallis et al. 1990). In contrast, the range of α,ω-dichlorinated alkanes that was used for growth of Pseudomonas sp. strain 273 was limited to the C9 and C10 substrates (Wischnak et al. 1998). Dehalogenase activity was demonstrated in a strain of Acinetobacter GJ70 that could degrade some α,ω-dichloroalkanes, and 1-bromo- and 1-iodopropane. Although 1,2-dibromoethane could be converted into 2-bromoethanol, this could not be used for growth plausibly due to the toxicity of bromoacetaldehyde that has already been noted, and the inability to use dihydroxyethane as growth substrate (Janssen et al. 1987). In a later study, the enzyme from this strain showed dehalogenase activity towards a wide range of substrates including halogenated alkanes, alkanols, and ethers (Janssen et al. 1988). A brief summary of the degradation of 1,2-dichloroethane by Xanthobacter autotrophicus strain GJ10 is given, since this strain has been used to delineate all stages of the metabolism and the appropriate enzymes have been demonstrated (Janssen et al. 1987) (Figure 7.62). The activity of a haloalkane dehalogenase initiates degradation and is discussed in this section, while the alkanol dehydrogenase, the aldehyde dehydrogenase, and the haloacetate dehalogenase are discussed subsequently. The enzyme responsible for dehalogenase activity has been purified from Xanthobacter autotrophicus strain GJ10, consists of a single polypeptide chain with a molecular mass of 36 kDa, and was able to dehalogenate chlorinated, and both brominated and iodinated alkanes

TABLE 7.2 Specific Activity (µmol bromide produced/mg protein/min) of Dehalogenase from Selected Species of Mycobacterium towards 1,2-Dibromoethane Taxon M. bovis BCG MU10 M. fortuitum MU8 M. triviale MU3 M. smegmatis CCM4622

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Activity

Taxon

Activity

99 76 61 49

M. avium MU1 M. phlei CCM 5639 M. parafortuitun MU2 M. chelonae

36 22 22 20

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Aliphatic Compounds

361

FIGURE 7.62 Degradation of 1,2-dichloroethane by Xanthobacter autotrophicus strain GJ10. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, pp. 1–74, Springer, 2002. With permission.)

Trp 75 N H

Phe 172

C2

Phe128 Cl

Cl

1

C1

2

H N O

O

Trp 125

H N

O

O

HN

Asp 260

N

H

O H

O

O

Asp 124

H N

His 289 Glu56 O

O

FIGURE 7.63 Mechanism of dehalogenation. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, pp. 1–74, Springer, 2002. With permission.)

(Keuning et al. 1985). Details of the mechanism have been explored using an ingenious method of producing crystal at different stages of the reaction (Verschuren et al. 1993). The overall reaction involves a catalytic triad at the active site: Asp124 binds to one of the carbon atoms, and hydrolysis with inversion is accomplished by cooperation of Asp260 and His289 with a molecule of water bound to Glu56 (Figure 7.63). Intermediate Epoxides

The conversion of vicinal haloalkanols into epoxides is of considerable interest and was apparently first recognized in a strain of Flavobacterium (Castro and Bartnicki 1968). The trans-epoxide from erythro-3-bromobutan-2-ol and the cis-epoxide from threo-3-bromobutan-2-ol were formed stereospecifically, and the epihalohydrins were able to react with halide to produce 2-hydroxy-1,3dihalobutanes (Bartnicki and Castro 1969). Haloalkanol dehalogenase activity has since been found in a number of different bacteria (Slater et al. 1997). The enantioselective halohydrin hydrogen lyase (alkanol dehalogenase) has been characterized from Corynebacterium sp. strain N-1074 (Nakamura et al. 1994a), as well as two different epoxide hydrolases that produce 3-chloropropane-1,2-diol from the same strain (Nakamura et al. 1994b). The haloalkanol dehalogenase has been characterized from Arthrobacter sp. strain AD2 that is able to utilize 3-chloro-1,2-propandiol for growth (van den Wijngaard et al. 1991). It has a molecular mass of 29 Da and consists of two equal subunits. The x-ray structure of the epoxide hydrolase from Agrobacterium radiobacter strain AD2 has been

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Environmental Degradation and Transformation of Organic Chemicals

carried out and provided details of its mode of action (Nardini et al. 1999). Haloalkanol dehalogenase from this strain is able to form epoxides from a number of vicinal chlorinated and brominated alkanols (van den Wijngaard et al. 1991), and is also able to carry out transhalogenation between epihalohydrins and halide ions. The activities of both haloalkanol dehalogenase (halohydrin hydrogen lyase) that catalyzes the formation of epoxides from alkanes with vicinal hydroxyl and halogen groups, and epoxide hydrolase that brings about hydrolysis of epoxyalkanes to diols are involved in a number of degradations that involve their sequential operation. a. In the degradation of epichlorohydrin (3-chloro-1,2-epoxyethane) (van den Wijngaard et al. 1989), the epoxide hydrolase produced 3-chloropropan-1,2-diol that was dehalogenated to glycidol by the dehydrogenase, followed by hydrolysis of the epoxide with the production of glycerol. b. The degradation of trihalogenated propanes by Agrobacterium radiobacter into which dehalogenases in plasmids from Rhodococcus sp. strain m15-3 were incorporated (Bosma et al. 1999) involved a sequence of steps, two of which involve the formation and hydrolysis of epoxides with the ultimate production of glycerol. Halogenated Alkanoates Halogenated alkanoic acids are produced from the corresponding aldehydes during the degradation of 1,2-dihaloethanes by Xanthobacter autotrophicus (Janssen et al. 1985) and the facultative methanotroph Ancylobacter aquaticus (van den Wijngaard et al. 1992), and are degraded by haloacetate dehalogenases. The aerobic degradation of halogenated alkanoic acids has been extensively investigated, and is generally carried out by halohydrolases that have variable specificity for the halogen, its position, and for different enantiomers. The dehalogenases have been grouped according to their reactions with 2-chloropropionate in carrying out hydrolysis with or without inversion—and the effect of sulfhydryl inhibitors (Fetzner and Lingens 1994). The enzymology of 2-haloalkanoate dehalogenases has been discussed in detail (Slater et al. 1997). The crystal structures of the complexes between the l-2-haloacid dehalogenase from Pseudomonas sp. strain YL and 2-chloroalkanoates have been determined (Li et al. 1998). They reveal that the hydrolytic inversion involves the Arg41 and Asp10 sites as electrophiles and nucleophiles, respectively, followed by interaction of the Asp10 ester with Ser118. The enantiomeric specific dehalogenase from Pseudomonas sp. strain DL-DEX is able to use both enantiomers of 2-haloalkanoic acids as substrates forming products with inversion of the configuration (Nardi-Del et al. 1997). The degradation of 2,2dichloropropionate involves dehalogenation to pyruvate but even here two different dehalogenases are synthesized (Allison et. al. 1983). Di- and trichloroacetate are degradable, though not apparently by the same groups of organisms. Mono- and dichloroacetate were effectively degraded by a strain of “Pseudomonas dehalogenans” that had only limited effect on trichloroacetate, whereas conversely, strains of Arthrobacter sp. readily degraded trichloroacetate—though not monochloroacetate (Jensen 1960). An organism that was able to grow with trichloroacetetae, though not with mono- or dichloroacetate was similar to Acinetobacter calcoaceticus on the basis of 16S rDNA (Yu and Welander 1995). The anaerobic degradation of halogenated alkanoic acids has, however, been much less exhaustively examined. Geobacter (Trichlorobacter) thiogenes was able to transform trichloroacetate to dichloroacetate by coupling the oxidation of acetate to CO2 with the reduction of sulfur to sulfide that carries out the dechlorination (De Wever et al. 2000). Halogenated Alkenoates The degradation of 2- and 3-chloroacrylate displays novel features. The degradation of 2-chloroacrylate by Burkholderia sp. strain WS is initiated by a 2-haloacrylate reductase with the formation of (S)-2-chloropropionate that undergoes dehalogenation to (R)-lactate (Kurata et al. 2005). The degradation of 1,3-dichloropropene by Pseudomonas pavonaceae (cichorii) strain 170 involves a series of steps by which trans-3-chloroacrylate is formed (Poelarends et al. 1998). This was degraded

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Aliphatic Compounds O

O

C

C

S N O C S

363

CCl4

Ligand

O C S

C S +CCl3 Lg

N Cu

O

Cu Cl

N O C S

COS S CCl3 Lg

Cl3CSH

Cl2

Cu

CO2

S CS2

Cl

FIGURE 7.64 Degradation of tetrachloromethane (CCl4) mediated by pyridine-2,6-dithiocarboxylate.

by dehalogenation to malonate semialdehyde (Poelarends et al. 2001), and it has been shown that this is accomplished by hydration of trans-3-chloroacrylate to an unstable halohydrin that collapses to form malonate semialdehyde (De Jong et al. 2004). The malonate semialdehyde decarboxylase (Poelarends et al. 2003) has been characterized. Glutathione-Mediated and Other Reactions Involving Nucleophilic Sulfur Although the reactions described above are formally nucleophilic displacements of the chlorine atoms by hydroxyl groups, a different mechanism clearly operates in the degradation of dichloromethane by Hyphomicrobium sp. The enzyme is glutathione-dependent (Stucki et al 1981; Kohler-Staub et al. 1986) and the reaction presumably involves at least two steps. Glutathionedependent reactions are also involved in the transformation and detoxification of alkenes including isoprene, and both cis- and trans-1,2-dichloroethene in Rhodococcus sp. strain AD45 (van Hylckama Vlieg et al. 1998). For cis-1,2-dichloroethene, the initially formed epoxide is transformed by reaction with glutathione followed by chemical reactions with the final production of glyoxal. Glutathione S-transferase and a glutathione dehydrogenase have been characterized in this strain (van Hylckama Vlieg et al. 1999). The degradation of tetrachloromethane by a strain of Pseudomonas sp. presents a number of exceptional features. Although 14CO2 was a major product from the metabolism of 14CCl4, a substantial part of the label was retained in nonvolatile water-soluble residues (Lewis and Crawford 1995). The nature of these was revealed by the isolation of adducts with cysteine and N,N′-dimethylethylenediamine, when the intermediates that are formally equivalent to COCl2 and CSCl2 were trapped—presumably formed by reaction of the substrate with water and a thiol, respectively. Further examination of this strain classified as Pseudomonas stutzeri strain KC has illuminated novel details of the mechanism. The metabolite pyridine-2,6-dithiocarboxylic acid (Lee et al. 1999) plays a key role in the degradation. Its copper complex produces trichloromethyl and thiyl radicals, and thence the formation of CO2, CS2, and COS (Figure 7.64) (Lewis et al. 2001). Monooxygenation Halogenated Alkanes A number of halogenated—including polyhalogenated—alkanes are degraded by hydroxylation mediated by MMO. 1. The soluble MMO from Methylococcus capsulatus (Bath) is able to oxidize chloro- and bromomethane—but not iodomethane—with the presumptive formation of formaldehyde (Colby et al. 1977). 2. The methane degrading Methylosinus trichosporium OB3b has been shown to degrade both methyl bromide and dibromomethane (Bartnicki and Castro 1994; Streger et al. 1999), and the propane-degrading Mycobacterium vaccae JOB5 can degrade methyl bromide (Streger et al. 1999). 3. A number of haloalkanes including dichloromethane, chloroform, 1,1-dichloroethane and 1,2-dichloroethane may be degraded by the soluble MMO system of Methylosinus trichosporium (Oldenhuis et al. 1989).

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Environmental Degradation and Transformation of Organic Chemicals

4. The metabolism of chloroform has been studied in several organisms (Hamamura et al. 1997), and from the inhibitory effect of acetylene it was concluded that a monooxygenase was involved: a. Cells of Methylosinus trichosporium strain OB3b grown with methanol and incubated with formate as electron donor degraded chloroform with the release of 2.1 mol of chloride per mole substrate and at substrate concentrations up to 38.6 µM. b. Cells of Pseudomonas butanovora grown with butane and incubated with butyrate as electron donor degraded chloroform with the release of 1.7 mol chloride per mole substrate and degradation was incomplete even at concentrations of 12.9 µM. Butane inhibited the degradation. This organism also partially degraded other chloroalkanes and chloroalkenes including chloroethene (vinyl chloride), 1,2-trans-dichloroethene and trichloroethene. c. Butane-grown cells of Mycobacterium vaccae strain JOB5 were able to degrade chloroform without addition of an electron donor. 5. The degradation of 1,2-dichloroethane by Pseudomonas sp. strain DCA1 was initiated not by hydrolysis, but by monooxygenation with the direct formation of 1,2-dichloroethanol that spontaneously decomposed to chloroacetaldehyde (Hage and Hartmans 1999). The cometabolism of halogenated methanes has been examined in Nitrosomonas europaea and may putatively be mediated by ammonia monooxygenase. a. Oxidation of methyl fluoride to formaldehyde has been demonstrated (Hyman et al. 1994), and of chloroalkanes at carbon atoms substituted with a single chlorine atom to the corresponding aldehyde (Rasche et al. 1991). b. Oxidation of a number of chloroalkanes and chloroalkenes including dichloromethane, chloroform, 1,1,2-trichloroethane, and 1,2,2-trichloroethene has been demonstrated (Vannelli et al. 1990). Although the rate of cometabolism of trihalomethanes increased with levels of bromine substitution, so also did the toxicity. Both factors must therefore be evaluated (Wahman et al. 2005). Halogenated Alkenes Although degradation of halogenated alkenes by direct displacement of halogen is not expected on purely chemical grounds, this reaction apparently occurs during the degradation of vinyl chloride by a strain of Pseudomonas sp. that carries out the direct hydrolysis to acetaldehyde followed by mineralization to CO 2 (Castro et al. 1992a). The degradation of 3-trans-3-chloroacrylate, however, involves dehalogenation by hydration to an unstable halohydrin (De Jong et al. 2004). The regulation of the synthesis of the soluble and particulate monooxygenase enzymes is illustrated by the degradation of trichloroethene by Methylosinus trichosporium strain OB3b. During copper limitation, the soluble monooxygenase is formed, but not during copper sufficiency when the particulate form is synthesized (Oldenhuis et al. 1989). Epoxidation is the first reaction in the degradation of chlorinated ethenes including chloroethene (vinyl chloride) (Coleman et al. 2002a; Coleman and Spain 2003; Danko et al. 2004) and cis-dichloroethene that is a recalcitrant intermediate in the anaerobic dechlorination of tetra- and trichloroethene (Coleman et al. 2002b). The epoxides formed by monooxygenation are toxic to cells and mechanisms must exist for their removal. Although evidence for the existence of epoxide hydrolases is well established in mammalian systems, this has not been so widely observed in bacteria. Some aspects of the pathways are summarized: 1. The degradation of trichloroethene (Figure 7.65) provides a good example of the monooxygenation of haloalkenes (Little et al. 1988). 2,2,2-Trichloroacetaldehyde is produced during oxidation of trichloroethene by several methanotrophs and undergoes a dismutation to form trichloroethanol and trichloroacetate (Newman and Wackett 1991). At least formally, this transformation is analogous to an NIH shift. The particulate enzyme contains

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Aliphatic Compounds

Hal Hal

C

C

365

O

Hal

Hal

H

Hal

C

C

Hal H

Hal Hal C Hal Hal Hal

CHO

CH−CO 2H

+

HO 2C−CHO

HCO 2 H + C

O

FIGURE 7.65 Generalized degradation of trihaloethenes by methanotrophs. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, pp. 1–74, Springer, 2002. With permission.)

copper, or both copper and iron, and the concentration of copper determines the catalytic activity of the enzyme (Sontoh and Semrau 1998). 2. The degradation of 1,2,3-trichloropropane by Agrobacterium radiobacter strain AD1 involves hydrolysis of an intermediate epichlorohydrin (3-chloroprop-1-ene) to the diol (Bosma et al. 1999, 2002). An enzyme from this strain has been modified from the use of epichlorohydrin that is its normal substrate to accept cis-1,2-dichloroethene with the release of chloride and the presumptive formation of glyoxal (Rui et al. 2004). 3. The aerobic degradation of chloroethene (vinyl chloride) by Mycobacterium aurum strain L1 proceeded by initial formation of an epoxide mediated by an alkene monooxygenase (Hartmans and de Bont 1992). This reaction has also been demonstrated to occur with Methylosinus trichosporium, even though subsequent reactions were purely chemical (Castro et al. 1992b). 4. The degradation of chloroethene by Mycobacterium sp. strain JS60 involved epoxidation followed by reaction with coenzyme M and loss of chloride in formation of the aldehyde. Subsequent steps involved reductive loss of coenzyme M and the production of acetate (Coleman and Spain 2003) and were analogous to those in Nocardioides sp. strain JS614 (Mattes et al. 2005). The reductive loss of coenzyme M is formally analogous to that involved in the degradation of isoprene by Rhodococcus sp. strain AD45 (van Hylckama Vlieg et al. 2000). The coenzyme M pathway is not limited to Gram-positive organisms and has also been demonstrated in Pseudomonas putida strain AJ and Ochrobactrum sp. strain TD (Danko et al. 2006). Induction of Monooxygenation by Non-Halogenated Substrates

An important observation is the transformation of chlorinated ethenes by monooxygenases induced by a range of “cosubstrates” including toluene, phenol, and methane or propane (references in Kageyama et al. 2005). Advantage has been taken of this for bioremediation of contaminated sites when both aromatic and chlorinated aliphatic contaminants are present. Factors determining the effectiveness of toluene as an inducer of monooxygenase have been examined (Leahy et al. 1996), although strains of Wautersia sp. can degrade trichloroethene in the absence of a cosubstrate (Kageyama et al. 2005). The degradation of trichloroethenes has also been examined by strains in which the relevant dioxygenase has been induced with a range of cosubstrates. These include (a) phenol or benzoate (Nelson et al. 1987), (b) toluene (Wackett and Gibson 1988), (c) isopropylbenzene (Pflugmacher et al. 1996), or (d) in a hybrid dioxygenase using elements of the toluene and biphenyl dioxygenase operons (Furukawa et al. 1994). Degradation seems to be limited to trichloroethene and cis-1,2-dichloroethene and the toxicity of the substrates is a limiting factor (Wackett and Gibson 1988). It seems possible that monooxygenation might be involved in view of the demonstrated existence of this activity in naphthalene dioxygenase (Gibson et al. 1995).

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Although monooxygenases can be involved in the degradation of toluene, there is a complex relation between toluene monooxygenase activity and the degradation of chlorinated hydrocarbons. Although toluene oxidation is accomplished by Pseudomonas mendocina strain KR1, P. putida strain F1, P. picketii strain PKO1, and Burkholderia cepacia strain G4 that possess toluene monooxygenase activities, the degradation of chloroform and 1,2-dichloroethane is carried out by only strains with the facility of toluene-4-monooxygenation—P. mendocina KR1 and Pseudomonas sp. strain ENVPC5 (McClay et al. 1996). In these strains, toluene oxidation could be induced by trichloroethene, which was subsequently degraded, whereas trichloroethene did not induce toluene oxidation in B. cepacia strain G4 or P. putida strain F1 (McClay et al. 1995). The degradation of trichloroethene by the three components of toluene 2-monooxygenase of B. cepacia involves initial formation of the epoxide followed by spontaneous decomposition to carbon monoxide, formate, and glycolate (Newman and Wackett 1997). In contrast to the degradation by MMO or cytochrome P450 monooxygenase, chloral hydrate was not formed. Reductive Reactions: Dehalogenation Including Dehalorespiration Cytochrome P450 reduction of polyhalogenated alkanes has been demonstrated. For example, cells of Pseudomonas putida G786 containing cytochrome P450cam genes on the CAM plasmid were able to carry out the reductive dehalogenation of a number of halogenated alkanes and for some substrates also oxidations. Illustrative examples include the following: 1. Selective reductive debromination of polyhalogenated methanes (Castro et al. 1985; Li and Wackett 1993). 2. Both aerobic and reductive pathways were suggested for the degradation of 1,1,2-trichloroethane (Castro and Belser 1990)—a dominant aerobic pathway to chloroacetate and glyoxylate, and simultaneously a minor reductive reaction, which must also involve an elimination reaction with the formation of chloroethene (Figure 7.66). 3. Under anaerobic conditions, various reactions can occur, and the following are illustrative: (a) trichlorofluoromethane → carbon monoxide; (b) hexachloroethane → tetrachloroethene; (c) 1,1,1-trichloro-2,2,2-trifluoroethane → 1,1-dichloro-2,2-difluoroethene (Hur et al. 1994). 4. When a plasmid containing genes encoding toluene dioxygenase was incorporated into this modified strain of P. putida, complete degradations that involved both reductive and oxidative steps were accomplished, for example, the degradation of pentachloroethane via trichloroethene to glyoxylate, and of 1,1-dichloro-2,2-difluoroethene to oxalate (Wackett et al. 1994). Anaerobic Dehalogenation of Polyhalogenated Alkenes Considerable effort has been devoted to the anaerobic transformation of polychlorinated C1 alkanes and C2 alkenes in view of their extensive use as industrial solvents and their identification as widely distributed groundwater contaminants. Early experiments, which showed that tetrachloroethene was transformed into chloroethene (Vogel and McCarty 1985) (Figure 7.67), aroused concern; though it has now been shown that complete dechlorination can be accomplished by some organisms.

Cl Cl

Cl CH2 CO2H CH

CH2Cl CH2

FIGURE 7.66

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CHCl

Biodegradation of trichloroethane by Pseudomonas putida.

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Aliphatic Compounds

367 Cl

H C

C

Cl

Cl

Cl C

C

Cl

Cl

H

Cl C Cl

H

Cl

H C

C Cl

H C

Cl

H

H

H

C Cl

CO2

H C

Cl

FIGURE 7.67

H

C

C Cl

Anaerobic dechlorination of tetrachloroethene.

The dechlorination of polychlorinated ethenes by a number of bacteria can be coupled to the synthesis of ATP and has been designated dehalorespiration (Holliger et al. 1999; Drzyzga and Gottschal 2002; Sun et al. 2002). Dehalorespiration has been demonstrated in Sulfurospirillum (Dehalospirillum) multivorans, Dehalobacter restrictus, Desulfuromonas chloroethenica, and some strains of Desulfitobacterium sp. (references in Holliger et al. 1999), and in Dehalococcoides ethenogenes that is capable of reductively dehalogenating tetrachloroethene to ethene (Magnuson et al. 2000; Maymó-Gatell et al. 1999). The electron donors were generally H2 or pyruvate, or methanol for Dehalococcoides ethenogenes. The chloroethene reductases contain corrinoid cofactors and [Fe–S] clusters, and the trichloroethene reductase from Dehalococcoides ethenogenes is able to debrominate substrates containing two, three, four, or five carbon atoms, albeit with decreasing ease (Magnuson et al. 2000). Further examples include the following dechlorinations: a. Polychlorinated ethenes to cis-1,2-dichloroethene, chloroethene (vinyl chloride), and ethene or ethane, and of chloroethene (vinyl chloride) to ethene (He et al. 2003; Müller et al. 2004). b. Tetrachloroethene to trans-1,2-dichloroethene (Griffin et al. 2004). c. Tetra- and trichloroethene at the expense of acetate or pyruvate and H2, to cis-dichloroethene by Geobacter lovleyi, which is a relative of Geobacter thiogenes that can couple growth from the dechlorination of trichloroacetate to dichloroacete (Sung et al. 2006). Among organisms that can dechlorinate tetrachloroethene, several types of reductase have been found. In Sulfurospirillum (Dehalospirillum) multivorans, Desulfitobacterium sp. strain PCE-S, Desulfitobacterium hafniense (frappieri) strain TCE-1, and Dehaloccoides ethenogenes, it consists of a peptide with a single corrinoid cofactor and two [Fe–S] centers. There is considerable interest in the dechlorination of tetrachloroethene and organisms that are able to do so include the following: 1. A strain of Sulfurospirillum (Dehalospirillum) multivorans transformed tetrachloroethene to trichloroethene and cis-1,2-dichloroethene (Neumann et al. 1994) using pyruvate as electron donor, and some properties of the dehalogenase have been reported (Neumann et al. 1995). One mol of the dehalogenase contained 1 mol of corrinoid, 9.8 mol of Fe, and 8.0 mol of acid-labile sulfur (Neumann et al. 1996), and the genes have been cloned and sequenced (Neumann et al. 1998). Although comparable values have been reported for the enzyme from Desulfitobacterium sp. strain PCE-S (Miller et al. 1997), the N-terminal sequence of the enzyme showed little similarity to that of Sulfurospirillum multivorans (Miller et al. 1998). The cardinal role of a corrinoid was proposed for both organisms (Neumann et al. 1996; Miller et al. 1997). A comparable reductive dehalogenation— though only as far as trichloroethene—is carried out by Desulfomonile tiedjei in which

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2.

3.

4.

5.

6.

7.

8.

9.

10.

Environmental Degradation and Transformation of Organic Chemicals

the reductase is a heme protein that is similar to that involved in dehalogenation of 3-chlorobenzoate (Townsend and Suflita 1996). Dehalococcoides ethenogenes is unusual among the dehalorespiring strains in its capability to reductively dehalogenate tetrachloroethene to ethene using methanol as substrate (Magnuson et al. 2000; Maymó-Gatell et al. 1999). It is also capable, though less readily, of reductively dehalogenating a few halogenated propenes including 3-chloropropene and 1,3-dichloropropene with the formation of propene (Magnuson et al. 2000). The membrane-bound dehalogenase from Dehalobacter restrictus contained (per mole of subunit) 1 mol of cobalamin, 0.6 mol cobalt, 7 mol iron, and 6 mol acid-labile sulfur. It carried out dechlorination of tetra- and trichloroethene to cis-1,2-dichloroethene, and had a substrate spectrum that included tetrachloromethane, hexachloroethane, and 1,1,1trichloro-2,2,2-trifluorarethane although the products from these were not apparently identified (Maillard et al. 2003). Two strains of Desulfuromonas michiganensis carried out the partial dechlorination of tetrachloroethene to cis-1,2-dichloroethene using acetate as the electron donor (Sung et al. 2003). The reduction of tetrachloroethene to cis-1,2-dichloroethene by the enteric organism Pantoea (Enterobacter) agglomerans may be noted (Sharma and McCarty 1996) as one of the few examples of the ability of Enterobacteriaceae to carry out reductive dechlorination. A nonfermentative organism putatively assigned to Desulfuromonas acetexigens reduced tetrachloroethene to cis-dichloroethene using acetate as electron donor (Krumholz et al. 1996), and a similar species D. chloroethenica used both tetra- and trichloroethene as electron acceptors with the production of cis-dichloroethene using acetate or pyruvate as electron donors (Krumholz 1997). Clostridium bifermentans can dechlorinate tetrachloroethene to cis-dichloroethene (Chang et al. 2000; Okeke et al. 2001). The enzyme has been purified and contains a corrinoid cofactor, although the complete nucleotide sequence of the gene encoding the reductase showed no homology with existing dehalogenases. The enzyme was most active with tetrachloroethene and trichloroethene, although lower rates were also observed with other chlorinated ethenes except chloroethene (vinyl chloride). Under methanogenic conditions, a strain of Methanosarcina sp. transformed tetrachloroethene to trichloroethene (Fathepure and Boyd 1988). In the presence of suitable electron donors such as methanol, complete reduction of tetrachloroethene to ethene may be achieved in spite of the fact that the dechlorination of vinyl chloride appeared to be the rate-limiting step (Freedman and Gossett 1989). Tetrachloroethene can be dechlorinated to trichloroethene as the sole product by the homoacetogen Sporomusa ovata using methanol as the electron donor, and cell extracts of other homoacetogens including Clostridium formicoaceticum and Acetobacterium woodii were able to carry this out using CO as electron donor (Terzenbach and Blaut 1994). Complete dechlorination of high concentrations of tetrachloroethene in the absence of methanogenesis has been achieved using methanol as electron donor (DiStefano et al. 1991).

There has been substantial interest in the complete dechlorination of chloroethene to ethene, since chloroethene is an undesirable product of partial dechlorination of polychlorinated ethenes. Strains of Dehalococcoides sp., and especially D. ethenogenes that is possibly unique in being able to carry out the dechlorination of tetrachloroethene to chloroethene and ethene (Magnuson et al. 2000; Maymó-Gatell et al. 1999), have attracted particular attention: a. It has been shown that the gene bvcA that is putatively involved in the reduction of chloroethene to ethene by Dehalococcoides sp. strain BAC1 was present in strains that could

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Aliphatic Compounds

369 BrCH2CH2Br BrCH

CH2 CH2

CHBr

HC

BrCH2CH2SO3H

CH

CH2 CH2

FIGURE 7.68 Reductive debromination by methanogenic bacteria. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, pp. 1–74, Springer, 2002. With permission.)

Cl

Cl

Cl

Cl

Cl

Cl

Cl Cl

Cl

Cl ;

Cl

Cl Cl

Cl Cl

Cl

FIGURE 7.69 Biotransformation of γ-hexachlorocyclohexene and γ-1,3,4,5,6-pentachlorocyclohexene. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, pp. 1–74, Springer, 2002. With permission.)

carry out chloroethene respiration, but was absent in those that could not (KrajmalnikBrown et al. 2004). b. The genes for the reductase designated vcrAB were detected at sites with established chloroethene contamination and it was suggested that this could serve as monitoring probe (Müller et al. 2004). Further examples include the debromination of 1,2-dibromoethane to ethene, and 1,2-dibromoethene to ethyne by methanogenic bacteria (Figure 7.68) (Belay and Daniels 1987). Details of these dehalogenations have emerged from studies with methanogens. The formation of ethene from 1,2-dichloroethane using hydrogen as electron donor has been demonstrated in cell extracts of Methanobacterium thermoautotrophicum DH and in Methanosarcina barkeri has been shown to involve cobalamin and F430 using Ti(III) as reductant (Holliger et al. 1992). Some additional comments on abiotic reactions are given in Chapter 1. Other Substrates 1. Reductive dehalogenation is one of the series of reactions involved in the degradation of γ-HCH (aaaeee) (Nagata et al. 1999): (a) initial elimination catalyzed by LinA, (b) hydrolysis by LinB, and (c) glutathione-mediated reductive loss of chloride from 2,5-dichlorohydroquinone catalyzed by LinD, which has already been noted (Figure 7.53) The last reaction also occurs during the degradation of pentachlorophenol (Xun et al. 1992). 2. Reductive dechlorination in combination with the elimination of chloride has been demonstrated in a strain of Clostridium rectum (Ohisa et al. 1982): γ-hexachlorocyclohexene formed 1,2,4-trichlorobenzene and γ-1,3,4,5,6-pentachlorocyclohexene formed 1,4dichlorobenzene (Figure 7.69). It was suggested that this reductive dechlorination is coupled to the synthesis of ATP, and this possibility has been clearly demonstrated during the dehalogenation of 3-chlorobenzoate coupled to the oxidation of formate in Desulfomonile tiedjei (Mohn and Tiedje 1991). Combined reduction and elimination has also been demonstrated in methanogenic cultures that transform 1,2-dibromoethane to ethene and 1,2dibromoethene to ethyne (Belay and Daniels 1987).

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3. Both the active enzyme, the heat-inactivated enzyme from Sulfurospirillum (Dehalospirillum) multivorans, and cyanocobalamin are capable of dehalogenating haloacetates (Neumann et al. 2002), and the rate of abiotic dehalogenation depends on the catalyst that is used. 4. An organism that is able to use methyl chloride as energy source and converting this into acetate has been isolated (Traunecker et al. 1991). 5. Cultures of a number of anaerobic bacteria are able to dechlorinate tetrachloromethane and Acetobacterium woodii formed dichloromethane as the final chlorinated metabolite by successive dechlorination, although CO2 was also produced by an unknown mechanism (Egli et al. 1988). 6. A strain of Clostridium sp. transformed 1,1,1-trichloroethane to 1,1-dichloroethane, and tetrachloromethane successively to trichloromethane and dichloromethane (Gälli and McCarthy 1989). 7. Although dichloromethane is a terminal metabolite in some transformations, an organism assigned to Dehalobacterium formicoaceticum is able to use this as a source of carbon and energy (Mägli et al. 1996). Dichloromethane was converted into methylenetetrahydrofolate from which formate is produced by oxidation and acetate by incorporation of CO2 catalyzed by CO dehydrogenase and acetyl-coenzyme A synthase. 8. Toxaphene is a complex mixture of compounds prepared by chlorinating camphene, and contains several hundred polychlorinated bornanes. After incubation with Sulfurospirillum (Dehalospirillum) multivorans only the hexa- (B6-923) and the heptachlorinated (B7-1001) remained. The nonachlorinated congener B9-1679 could be transformed by preferential dechlorination of the gem-dichloro groups with formation of the heptachlorinated B7-1001 (Ruppe et al. 2003). Further examination showed that this congener could be dechlorinated to penta- and hexachlorinated bornanes (Ruppe et al. 2004).

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Traunecker J, A Preub, G Diekert (1991) Isolation and characterization of a methyl chloride utilizing, strictly anaerobic bacterium. Arch Microbiol 156: 416–421. Tu CM (1976) Utilization and degradation of lindane by soil microorganisms. Arch Microbiol 108: 259–263. van den Wijngaard AJ, PTW Reuvenkamp, DB Janssen (1991) Purification and characterization of haloalcohol dehalogenase from Arthrobacter sp. strain AD2. J Bacteriol 173: 124–129. van den Wijngaard, AJ, DB Janssen, B Withold (1989) Degradation of epichlorohydrin and halohydrins by bacterial cultures isolated from freshwater sediment. J Gen Microbiol 135: 2199–2208. van den Wijngaard, AJ, J Prins, AJAC Smal, DB Janssen (1993) Degradation of 2-chloroethylvinyl ether by Ancylobacter aquaticus AD25 and AD27. Appl Environ Microbiol 59: 2777–2783. van den Wijngaard, AJ, KWHJ van der Kamp, J van der Ploeg, F Pries, B Kazemier, DB Janssen (1992) Degradation of 1,2-dichloroethane by Ancyclobacter aquaticus and other facultative methylotrophs. Appl Environ Microbiol 58: 976–983. van der Ploeg, M Willemsen, G van Hall, DB Janssen (1995) Adaptation of Xanthobacter autotrophicus GJ10 to bromoacetate due to activation and mobilization of the haloacetate dehalogenase gene by insertion element IS1247. J Bacteriol 177: 1348–1356. van Hylckama Vlieg JET, H Leemhuis, JHL Spelberg, DB Janssen (2000) Characterization of the gene cluster involved in isoprene metabolism in Rhodococcus sp. strain AD45. J Bacteriol 182: 1956–1963. van Hylckama Vlieg JET, J Kingma, AJ van den Wijngaard, DB Janssen (1998) A gluathione S-transferase with activity towards cis-1,2-dichloroepoxyethane is involved in isoprene utilization by Rhodococcus strain AD 45. Appl Environ Microbiol 64: 2800–2805. van Hylckama Vlieg JET, J Kingma, W Kruizinga, DB Janssen (1999) Purification of a glutathione S-transferase and a glutathione conjugate-specific dehydrogenase is involved in isoprene metabolism by Rhodococcus sp. strain AD 45. J Bacteriol 181: 2094–2101. Vannelli T, M Logan, DM Arciero, AB Hooper (1990) Degradation of halogenated aliphatic compounds by the ammonia-oxidizing bacterium Nitrosomonas europaea. Appl Environ Microbiol 56: 1169–1171. Vannelli T, M Messmer, A Studer, S Vuilleumier, T Leisinger (1999) A corrinoid-dependent catabolic pathway for growth of a Methylobacterium strain with chloromethane. Proc Natl Acad Sci USA 96: 4615–4620. Verschuren KHG, F Seljée, HJ Rozeboom, KH Kalk, BW Dijkstra (1993) Crystallographic analysis of the catalytic mechanism of haloalkane dehalogenase. Nature 363: 693–698. Vogel TM, PL McCarty (1985) Biotransformation of tetrachloroethylene to trichloroethylene, dichloroethylene, vinyl chloride, and carbon dioxide under methanogenic conditions. Appl Environ Microbiol 49: 1080–1083. Wackett LP, DT Gibson (1988) Degradation of trichloroethylene by toluene dioxygenase in whole-cell studies with Pseudomonas putida F1. Appl Environ Microbiol 54: 1703–1708. Wackett LP, MJ Sadowsky, LM Newman, H-G Hur, S Li (1994) Metabolism of polyhalogenated compounds by a genetically engineered bacterium. Nature 368: 627–629. Wahman DG, LE Katz, GE Speitel (2005) Cometabolism of trihalomethanes by Nitrosomonas europaea. Appl Environ Microbiol 71: 7980–7986. Wedemeyer G (1967) Dechlorination of 1,1,1-trichloro-2,2-bis[p-chlorophenyl]ethane by Aerobacter aerogenes. I. Metabolic products. Appl Microbiol 15: 569–574. Willett KL, EM Ulrich, RA Hites (1998) Differential toxicity and environmental fates of hexachlorocyclohexane isomers. Environ Sci Technol 32: 2197–2207. Wischnak C, FE Löffler, J Li, JW Urbance, R Müller (1998) Pseudomonas sp. strain 273, an aerobic α,ωdichloroalkane-degrading bacterium. Appl Environ Microbiol 64: 3507–3511. Woodall CA, KL Warner, RS Oremland, JC Murrell, IR McDonald (2001) Identification of methyl halideutilizing genes in the methyl bromide-utilizing bacterial strain IMB-1 suggests a high degree of conservation of methyl halide-specific genes in Gram-negative bacteria. Appl Environ Microbiol 67: 1959–1963. Xun L, E Topp, CS Orser (1992) Diverse substrate range of a Flavobacterium pentachlorophenol hydroxylase and reaction stoichiometries. J Bacteriol 174: 2898–2902. Yu P, T Welander (1995) Growth of an aerobic bacterium with trichloroacetic acid as the sole source of energy and carbon. Appl Microbiol Biotechnol 42: 769–774.

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PART 4 FLUORINATED ALIPHATIC COMPOUNDS There has been increase in the application of these compounds since the synthesis of fluorinated alkanes and related compounds in the 1930s. These include fluorinated hydrocarbons that were formerly used as propellants, polymerized tetrafluoroethene, and the polyfluorinated C4 –C8 carboxylates and sulfonates. All of them are notable for their inertness under normal conditions. Aromatic fluorinated compounds are discussed in Chapter 9, Part 3.

ALKANES AND ALKENES The biodegradation of hydrochlorofluorocarbons and hydrofluorocarbons has attracted considerable attention on account of their presumptive adverse effect on ozone depletion and climate alteration (references in Fabian and Singh 1999). Valuable background on chlorofluorocarbons has been given (Elliott 1994) and on alternatives to them (Rao 1994). The perfluorinated compounds including difluoromethane (HFC-32), trifluoroethane (HFC-143a), tetrafluoroethane (HFC-134a), hexafluoropropane (HFC-236ea), and heptafluoropropane (HFC-227ea) are, however, of potential interest as environmentally acceptable (Harnisch 1999). A list of acronyms is given in Table 7.3 (Midgley and McCulloch 1999). Further details of their degradation are given in a review (Neilson and Allard 2002). Fluorohydrocarbons Fluoromethane has been used as a selective inhibitor of ammonium oxidation and nitrificationlinked synthesis of N2O in Nitrosomonas europaea (Miller et al.1993), while difluoromethane has been proposed as a reversible inhibitor of methanotrophs (Miller et al. 1998). Monooxygenation is distributed among a variety of bacteria and several have been examined for their potential to degrade fluorinated alkanes: 1. Ammonia monooxygenase in Nitrosomonas europaea is able to oxidize fluoromethane to formaldehyde (Hyman et al. 1994). 2. The methanotroph Methylosinus trichosporium strain OB3b that produces the soluble MMO system consisting of a 40 kDa NADH oxidoreductase, a 245 kDa hydroxylase, and a 16 kDa protein termed component B has a low substrate specificity (Sullivan et al. 1998). It has been shown to metabolize trifluoroethene to glyoxylate, difluoroacetate, and the rearranged product trifluoroacetaldehyde (Fox et al. 1990). The last reaction is analogous to the formation of trichloroacetaldehyde from trichloroethene by the same strain (Oldenhuis et al. 1989).

TABLE 7.3 Acronyms for chlorofluorocarbons (CFCs), hydrochlorofluorocarbons (HCFCs), and hydrofluorocarbons (HFCs) Acronym CFC-11 CFC-12 CFC-113 CFC-114 CFC-115

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Structure CCl3F CCl2F2 CCl2F–CClF2 CClF2–CClF2 CClF3–CF3

Acronym HCFC-123 HCFC-22 HCFC-141b HCFC-142b HCFC-124 HCFC-225ca HCFC-225cb

Structure CHCl2–CF3 CHClF2 CH3–CCl2F CH3–CClF2 CHClF–CF3 CHCl2–CF2–CF3 CHFCl–CF2–CF3Cl

Acronym HFC-134a HFC-227ea HFC-152a HFC-143a HFC-125 HFC-32 HFC-23 HFC-245ca

Structure CH2F–CF3 CF3–CHF–CF3 CH3–CHF2 CH3–CF3 CHF2–CF3 CH2F2 CHF3 CHF2–CF2–CH2F

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379

3. This strain and Mycobacterium vaccae strain JOB 5 that produces propane monooxygenase has been used to examine the degradation of a number of hydrochlorofluorocarbons and hydrofluorocarbons (Streger et al. 1999). It was shown that during complete degradation by M. trichosporium and M. vaccae, this was accompanied by release of fluoride. It is worth noting, however, that the methylotrophs IMB-1 (Schaefer and Oremland 1999) and CC495 (Coulter et al. 1999) are able to oxidize chloromethane, bromomethane, and iodomethane, though not fluoromethane. On the basis of their 16S rRNA sequences, these organisms are related to those classified as Pseudaminobacter sp., and more distantly related to the nitrogen-fixing rhizobia (Coulter et al. 1999). Chlorofluorocarbons and Hydrochlorofluorocarbons In general, halogen is last from organic substrates in the order I > Br > Cl > F. Considerable effort has been directed to degradation and transformation by methylotrophic bacteria. As a generalization, the HCFCs are more readily degraded that the corresponding compounds lack hydrogen. The soluble MMO from Methylosinus trichosporium OB3b that has a wide substrate spectrum is able to oxidize some hydrochlorofluoroethanes including 1,1,2-trichloro-2-2-fluororethane and 1,1,2-trifluoroethane, though neither trichlorofluoromethane (CFC-11) nor any hydrochlorofluoroethane with three fluorine substituents on the same carbon atom were oxidized (DeFlaun et al. 1992). Although quantitative loss of fluoride and chloride was shown for dichlorofluoromethane, none of the organic products of oxidation were identified for the others. Reductive dechlorination of fluorinated substrates has been observed for both methanogens and sulfate-reducing bacteria. 1. The methanogen Methanosarcina barkeri is able to transform trichlorofluoromethane (CFC-11) by successive loss of chlorine to produce chlorofluoromethane (Figure 7.70a) (Krone and Thauer 1992). A similar transformation has been demonstrated in the presence of sulfate and butyrate with a mixed culture containing putatively Desulfovibrio baarsii and Desulfobacter postgatei (Sonier et al. 1994). 2. In anaerobic microcosms, 1,1,2-trichloro-1,2,2-trifluoroethane (CFC-113) was transformed by successive reductive dechlorination to 1,2-dichloro-1,2,2-trifluoroethane (HCFC-123a), and under methanogenic conditions to 1-chloro-1,2,2-trifluoroethane (HCFC-133) and 1-chloro-1,1,2-trifluoroethane (HCFC-133b) without evidence for the reductive replacement of fluorine (Figure 7.70b) (Lesage et al. 1992). 3. 1,1,1-Trifluoro-2,2-dichloroethane (HCFC-123) was recalcitrant in aerobic soils, but underwent reductive dechlorination anaerobically to produce 1,1,1-trifluoro-2-chloroethane (Oremland et al. 1996). Cytochrome P450cam is a monooxygenase that can be induced in Pseudomonas putida G786 by growth on camphor, and is responsible for the introduction of a hydroxyl group at C5 as the first step in the degradation of camphor. This enzyme is also able to carry out nonphysiological reductive dehalogenation (Castro et al.1985), and it has been shown (Li and Wackett 1993) that the electrons for reduction can be supplied by putidaredoxin. These alternatives are summarized: Monooxygenation Reductive dehalogenation

R–H + O2 + 2H+ + 2e → R–OH + H2O R–Hal + 2H+ + 2e → R–H + H–Hal

The production of carbon monoxide from trichlorofluoromethane catalyzed by cytochrome P450cam proceeded through intermediate formation of the dichlorofluorocarbene (Li and Wackett 1993) (Figure 7.70c). Other reactions included β-elimination from 1,1,1-trichloro-2,2,2-trifluorethane (Figure 7.70c). Pseudomonas putida strain G786 (pGH-2) was constructed to contain both the

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Cl (a) Cl Cl

C

F

Cl (b) Cl F

C

C

Cl (c) Cl Cl

C

F

Cl Cl Cl

C

C

Cl H H

Cl F F

C

F

Cl F H

C

F

Cl Cl

F

C

C

Cl

F C

F + H F

C

F F

H

CO

H

F

F F

[ Cl2 CF ]

H

Cl F H

Cl C

F C

C

F

Cl F

C

C F

Cl

F

FIGURE 7.70 (a) Metabolism of trichlorofluoromethane by Methanosarcina barkeri, (b) transformation of 1,1,2-trichloro-1,2,2-trifluoroethane (CFC-113) under anaerobic conditions, (c) dehalogenation by cytochrome P450cam of trichlorofluoromethane and 1,1,1-trichloro-2,2,2-trifluoroethane. Cl

Cl Cl Cl

C

C

F F

(a)

F

Cl C

C

Cl

Cl

Cl +

Cl

F

C H

C

F F

(b) HO2C−CO2H

(a) anaerobic,

(b) aerobic

FIGURE 7.71 Combined dehalogenation of 1,1,1,2-tetrachloro-2,2-difluoroethane by cytochrome P450cam and oxidation with toluene 2,3-dioxygenase.

cytochrome P450cam genes on the CAM plasmid and the tod C1, tod C2, tod b, and tod A genes of toluene dioxygenase. Toluene dioxygenase was constitutively expressed and cytochrome P450cam after induction by camphor (Wackett et al. 1994). Under anaerobic conditions 1,1,1,2-tetrachloro2,2-difluoroethane was dehalogenated to 1,1-chloro-2,2-difluoroethene that could be oxidized by the dioxygenase under aerobic conditions to oxalate (Figure 7.71) (Hyman et al. 1994). Carboxylic Acids Fluorinated aliphatic carboxylic acids are produced by some plants including species of Dichapetalum (Gribble 2002), and several mechanisms have been suggested for the biosynthesis of fluorinated metabolites (Harper et al. 2003). The mechanism of fluoroacetate toxicity in mammals has been extensively examined and was originally thought to involve simply initial synthesis of fluorocitrate that inhibits aconitase and thereby functioning the TCA cycle. Walsh has extensively reinvestigated the problem, and revealed both the complexity of the mechanism of inhibition and the stereospecificity of the formation of fluorocitrate from fluoroacetate (Walsh 1982).

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Aliphatic Compounds CO 2 H

381 F

CO 2 H F

F

O

HO 2C

CO 2 H

OH NO 2

O

CO 2 H

CH(F)CO 2 H CO 2 H

OH F O

CO 2 H CO 2 H

FIGURE 7.72

+

Degradation of 2-fluoro-4-nitrobenzoate by Nocardia erythropolis.

Fluoroacetate As might be expected, bacteria have been isolated from the plants that produce fluoroacetate, and these include an unidentified Pseudomonas sp. (Goldman 1965), a strain of Burkholderia (Pseudomonas) cepacia from Dichapetalum cymosum (Meyer et al. 1990), and a strain of Moraxella sp. (Kawasaki et al. 1981). In addition, fluoroacetate is an unusual product of microbial metabolism: a. It is a terminal metabolite formed during the metabolism of 2-fluoro-4-nitrobenzoate by Nocardia erythropolis (Figure 7.72) (Cain et al. 1968). b. It is formed together with 4-fluorothreonine during the late-stage growth of Streptomyces cattleya on a defined medium in the presence of fluoride (Reid et al. 1995). It has been shown (Hamilton et al. 1998) that glycine is an effective precursor of both fluoroacetate and 4-fluorothreonine, and that glycine is metabolized via N5,N10-methylenetetrahydrofol ate to serine and thence to pyruvate. However, details of the mechanism for the incorporation of fluoride remain incompletely resolved. The metabolism of fluoroacetate results in the production of fluoride and glycolate, and the use of H218O showed that the oxygen atom was introduced from water by a hydroxylase (Goldman 1966). The enzyme has been purified from the unidentified pseudomonad (Goldman 1965), and is specific for fluoroacetate: it is inactive towards 2- and 3-fluoropropionate, di- and trifluoroacetates, and fluorobenzoates. It is rapidly inhibited by 4-chloromercuriphenylsulfonate and slowly by N-methylmaleimide, and this suggests the involvement of an active thiol group in the enzyme. Strains of Moraxella sp. have also been shown to assimilate fluoroacetate using plasmid-determined dehalogenase activities (Kawasaki et al. 1981). One of these (strain H-1) was active towards both fluoroacetate and chloroacetate, whereas the other (strain H-2) was active only towards chloroacetate. Further investigation (Au and Walsh 1984) of the haloacetate hydrolase H-1 from Pseudomonas sp. strain A used 1H NMR of the (−)-α-methoxyl-α-(trifluoromethyl) phenylacetic acid ester of the phenacyl ester of the glycolate produced from (S)-2-[2H1]fluoroacetate. This established that the major metabolite was the (R)-enantiomer of 2-[2H1]glycolate and that the reaction proceeded with inversion of the configuration at C-2. The mechanism for the hydrolysis of fluoroacetate by Moraxella sp. strain B has been examined and confirmed the involvement of Asp105 as nucleophile with inversion of the configuration at C2 (Liu et al. 1998). Di- and Trifluoroacetate The soluble MMO from Methylsosinus trichosporium OB3 produced difluoroacetate as one of the main products from the oxidation of trifluoroethene and a low yield of trifluoroacetaldehyde by rearrangement (Fox et al. 1990).

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Chlorodifluoroacetic acid has been identified in rain and snow samples (Martin et al. 2000) and may plausibly be an atmospheric degradation product of 1,1,2-trichloro-1,2,2-trifluoroethane. Trifluoroacetate has been found in a wide range of environmental samples (Key et al. 1997), and attention has been directed to its origin, its toxicity, and its recalcitrance. Compared with chloroacetates, however, it is only mildly phytotoxic to algae (Berends et al. 1999) and higher plants (Boutonnet et al. 1999). It is worth noting its formation during the photochemical degradation of 3-trifluoromethyl-4-nitrophenol (Ellis and Mabury 2000). Concern has been expressed over its biodegradability and conflicting results over its recalcitrance under anaerobic conditions have been reported. This has been resolved by the results of a chemostat study using a mixed ethanol-degrading culture (Kim et al. 2000). Clear evidence for anaerobic degradation was produced on the basis of fluoride release and formation of acetic acid and methane. Perfluoroalkyl Carboxylates and Sulfonates There has been increasing interest in perfluoroalkyl sulfonates and carboxylates (Key et al. 1997). The sulfonates are valuable surface protectors and surfactants under extreme conditions, and are components of fire fighting foams (Moody and Field 2000). Perfluorooctane carboxylates and sulfonates have a global distribution in wildlife and humans (Houde et al. 2006). Low levels of perfluoro compounds have been carefully quantified in samples of human sera (Hansen et al. 2001) and determination of perfluorinated surfactants in surface water samples has been described. Two independent analytical techniques—liquid chromatography/tandem mass spectrometry and 19F NMR were used—and a summary of analytical methods was included (Moody et al. 2001). References to toxicological studies in rats and epidemiological studies in man have been given (Hansen et al. 2001), and the biochemical toxicology of the related perfluorooctanoic acid has been discussed (DePierre 2002). Concern with their apparent persistence has motivated investigations of their potential sources including fluorotelomer alcohols. A strain of Pseudomonas sp. strain D2 was used to evaluate the degradation of a range of fluorinated sulfones (Key et al. 1998). Although the fully fluorinated perfluorooctane sulfonate was resistant to defluorination by an aerobic bacterium, the presence of hydrogen substituents in 1H,1H,2H,2H-perfluorooctane sulfonate made it possible for the organism to carry out partial defluorination. For growth and defluorination of difluoromethane sulfonate, acetate or glucose were supplied as the source of carbon, ammonium as the source of nitrogen, and the substrate as the source of sulfur. It should be noted that perfluoro analogs that lacked hydrogen substituents did not support growth of this strain. These considerations apply also to the fluorotelomer alcohol CF3(CF2)7–CH2CH2OH that was degraded in a mixed culture obtained by enrichment with ethanol. Terminal dehydrogenation followed by elimination of fluoride, hydration and further loss of fluoride produced perfluorooctanoate (Dinglasan et al. 2004). CF3(CF2)7CH2–CH2OH → CF3(CF2)7CH2CO2H → CF3(CF2)6 –CF=CHCO2H → CF3(CF2)6COCH2CO2H → CF3(CF2)6CO2H An exhaustive study with activated municipal sludge using the telomer alcohol labeled with 14C at the terminal CF2 established the formation of a range of products that involved fission of several C–F bonds (Wang et al. 2005). Perfluorooctanoate is also formed during simulated atmospheric reactions of the fluorotelomer alcohol (Ellis et al. 2004) and it is suggested that all these reactions are the possible sources of perfluorocarboxylic acids.

REFERENCES Au KG, CT Walsh (1984) Stereochemical studies on a plasmid-coded fluoroacetate halohydrolase. Bioorg Chem 12: 197–295. Berends AG, JC Bouttonet, CG de Rooij, RS Thompson (1999) Toxicity of trifluoroacetate to aquatic organisms. Environ Toxicol Chem 18: 1053–1059.

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383

Boutonnet JC et al. (1999) Environmental risk assessment of trifluoroacetic acid. Human Ecol Risk Assess 5: 59–124. Cain RB, EK Tranter, JA Darrah (1968) The utilization of some halogenated aromatic acids by Nocardia. Oxidation and metabolism. Biochem J 106: 211–227. Castro CE, RS Wade, NO Belser (1985) Biodehalogenation reactions of cytochrome P-450 with polyhalomethanes. Biochemistry 24: 204–210. Coulter C, JTG Hamilton, WC McRoberts, L Kulakov, MJ Larkin, DB Harper (1999) Halomethane : bisulfide/halide ion methyltransferase, an unusual corrinoid enzyme of environmental significance isolated from an aerobic methylotroph using chloromethane as the sole carbon source. Appl Environ Microbiol 65: 4301–4312. DeFlaun ME, BD Ensley, RJ Steffan (1992) Biological oxidation of hydrochlorofluorocarbons (HCFCs) by a methanotrophic bacterium. Biotechnology 10: 1576–1578. DePierre JW (2002) Effects on rodents of perfluorofatty acids. Handbook Environ Chem 3N: 203–248. Dinglasan MJA, Y Ye, EA Edwards, SA Mabury (2004) Fluorotelomer alcohol biodegradation yields polyand perfluorinated acids. Environ Sci Technol 38: 2857–2864. Elliott AJ (1994) Chlorofluorocarbons. In Organofluorine Chemistry Principles and Commercial Applications (Eds RE Banks, BE Smart, and JC Tatlow), pp. 145–157. Plenum Press, New York. Ellis DA, JW Martin, AO De Silva, SA Mabury, MD Hurley, MPS Andersen, TJ Wallington (2004) Degradation of fluorotelomer alcohols: a likely atmospheric source of perfluorinated carboxylic acids. Environ Sci Technol 38: 3316–3321. Ellis DA, SA Mabury (2000) The aqueous photolysis of TFM and related trifluoromethylphenols. An alternate source of trifluoroacetic acid in the environment. Environ Sci Technol 34: 632–637. Fabian P, ON Singh (Eds) (1999) Reactive halogen compounds in the atmosphere. Handb Environ Chem 4E. Fox BG, JG Borneman, LP Wackett, JD Lipscomb (1990) Halolkene oxidation by the soluble methane monooxygenase from Methylosinus trichosporium OB3b: mechanistic and environmental implications. Biochemistry 29: 6419–6427. Goldman P (1965) The enzymatic cleavage of the carbon-fluorine bond in fluoroacetate. J Biol Chem 240: 3434–3438. Goldman P (1966) Carbon-fluorine bond cleavage II Studies on the mechanism of the defluorination of fluoroacetate. J Biol Chem 241: 5557–5559. Gribble GW (2002) Naturally occurring organofluorines. Handbook Environ Chem 3N: 121–136. Hamilton JTG, CD Murphy, MR Amin, D O′Hagan, DB Harper (1998) Exploring the biosynthetic origin of fluoroacetate and 4-fluorothreonine in Streptomyces cattleya. J Chem Soc Perkin Trans 1: 759–767. Hansen KJ, LA Clemen, ME Ellefson, HO Johnson (2001) Compound-specific quantitative characterization of organic fluorochemicals in biological matrices. Environ Sci Technol 35: 766–770. Harnisch J (1999) Reactive Fluorine Compounds. Handbook Environ Chem 4E: 81–111. Harper DB, D O′Hagan, CD Murphy (2003) Fluorinated natural products: occurrence and biosynthesis. Handbook Environ Chem 3P: 141–169. Houde M, JW Martin, RJ Letcher, KR Solomon, DCG Muir (2006) Biological monitoring of perflkuoroalkyl sybstances: a review. Environ Sci Technol 40: 3463–3473. Hyman MR, CL Page, DJ Arp (1994) Oxidation of methyl fluoride and dimethyl ether by ammonia monooxygenase in Nitrosomonas europaea. Appl Environ Microbiol 60: 3033–3035. Kawasaki H, N Tone, K Tonomura (1981) Plasmid-determined dehalogenation of haloacetates in Moraxella species. Agric Biol Chem 45: 29–34. Key BD, RD Howell, CS Criddle (1997) Fluorinated organics in the biosphere. Environ Sci Technol 31: 2445–2454. Key BD, RD Howell, CS Criddle (1998) Defluorination of organofluorine sulfur compounds by Pseudomonas sp. strain D2. Environ Sci Technol 32: 2283–2287. Kim BR, MT Suidan, TJ Wallington, and X Du (2000) Biodegradability of trifluoroacetic acid. Environ Eng Sci 17: 337–342. Krone UE, RK Thauer (1992) Dehalogenation of trichlorofluoromethane (CFC-11) by Methanosarcina barkeri. FEMS Microbiol Lett 90: 201–204. Lesage S, S Brown, KH Hosler (1992) Degradation of chlorofluorocarbon-113 under anaerobic conditions. Chemosphere 24: 1225–1243. Li S, LP Wackett (1993) Reductive dehalogenation by cytochrome P450 CAM: substrate binding and catalysis. Biochemistry 32: 9355–9361.

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Liu J-Q, T Kurihara, S Ichiyama, M Miyagi, S Tsunasawa, H Kawasaki, K Soda, N Esaki (1998) Reaction mechanism of fluoroacetate dehalogenase from Moraxella sp. B. J Biol Chem 273: 30897–30902. Martin JW, J Franklin, ML Hanson, KR Solomon, SA Mabury, DA Ellis, BF Scott, and DCG Muir (2000) Detection of chlorodifluoroacetic acid in precipitation: a possible product of fluorocarbon degradation. Environ Sci Technol 34: 274–281. Meyer JJM, N Grobbelaar, PL Steyn (1990) Fluoroacetate-metabolizing pseudomonad isolated from Dichapetalum cymosum. Appl Environ Microbiol 56: 2152–2155. Midgley PM, A McCulloch (1999) Properties and applications of industrial halocarbons. In Reactive Halogen Compounds in the Atmosphere (Eds P Fabian and ON Singh). Springer, Berlin. Miller LG, C Sasson, RS Oremland (1998) Difluoromethane, a new and improved inhibitor of methanotrophy. Appl Environ Microbiol 64: 4357–4362. Miller LG, MD Coutlakis, RS Oremland, BB Ward (1993) Selective inhibition of ammonium oxidation and nitrification-linked N2O formation by methyl fluoride and dimethyl ether. Appl Environ Microbiol 59: 2457–2464. Moody CA, JA Field (2000) Perfluorinated surfactants and environmental implications of their use in firefighting foams. Environ Sci Technol 34: 3864–3870. Moody CA, WC Kwan, JW Martin, DCG Muir, SA Mabury (2001) Determination of perfluorinated surfactants in surface water samples by two independent analytical techniques: liquid chromatography/ tandem mass spectrometry and 19F NMR. Anal Chem 73: 2200–2206. Neilson AH, A-S Allard (2002) Degradation and transformation of organic fluorine compounds. Handbook Environ Chem 3N: 138–202. Oldenhuis R, RLJM Vink, DB Janssen, B Witholt (1989) Degradation of chlorinated aliphatic hydrocarbons by Methylosinus trichosporium OB3b expressing soluble methane monooxygenase. Appl Environ Microbiol 55: 2819–2826. Oremland RS, DJ Lonrergan, CW Culbertson, DR Lovley (1996) Microbial degradation of hydrochlorofluorocarbons (CHCl2F and CHCl2CF3) in soils and sediments. Appl Environ Microbiol 62: 1818–1821. Rao VNM (1994) Alternatives to chlorofluorocarbons (CFCs). In Organofluorine Chemistry Principles and Commercial Applications (Eds RE Banks, BE Smart, and JC Tatlow), pp. 159–175. Plenum Press, New York. Reid KA, JTG Hamilton, RD Bowden, DO′Hagan, L Dasaradhi, MR Amin, DB Harper (1995) Biosynthesis of fluorinated secondary metabolites by Streptomyces cattleya. Microbiology (UK) 141: 1385–1393. Schaefer JK, RS Oremland (1999) Oxidation of methyl halides by the facultative methylotroph strain IMB-1. Appl Environ Microbiol 65: 5035–5041. Sonier DN, NLDuran, GB Smith (1994) Dechlorination of trichlorofluoromethane (CFC-11) by sulfatereducing bacteria from an aquifer contaminated with halogenated aliphatic compounds. Appl Environ Microbiol 60: 4567–4572. Streger SH, CW Condee, AP Togna, MF Deflaun (1999) Degradation of halohydrocarbons and brominated compounds by methane- and propane-oxidizing bacteria. Environ Sci Technol 33: 4477–4482. Sullivan JP, D Dickinson, HA Chase (1998) Methanotrophs, Methylosinus trichosporíum OB3b, sMMO, and their application to bioremediation. Crit Rev Microbiol 24: 335–373. Wackett LP, M J Sadowsky, LM Newman, H-G Hur, S Li (1994) Metabolism of polyhalogenated compounds by a genetically engineered bacterium. Nature (London) 368: 627–629. Walsh C (1982) Fluorinated substrate analogs: routes of metabolism and selective toxicity. Adv Enzymol 55: 187–288. Wang N, B Szostek, RC Buck, PW Folsom, LM Sulecki, V Capka, , MR Berti, JT Gannon (2005) Fluorotelomer alcohol biodegradation—direct evidence that perfluorinated carbon chains breakdown. Environ Sci Technol 39: 7516–7628.

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Before the advent of the petrochemical industry carbocyclic aromatic compounds, such as naphthalene, phenol, and pyridine, provided the source of many important industrial chemicals including dyestuffs, while the monocyclic compounds continue to play an important role as fuels and starting materials.

PART 1 MONOCYCLIC AROMATIC HYDROCARBONS INTRODUCTION The degradation of aromatic compounds including hydrocarbons and phenols has attracted interest over many years, for several reasons: 1. They are components of unrefined oil, and there has been serious concern over the hazard associated with their discharge into the marine environment after accidents at sea. 2. A number of the polycyclic representatives have been shown to be human procarcinogens that require metabolic activation. 3. There has been increased concern over air pollution as a result of their presence in the atmosphere from incomplete combustion. 4. The corresponding phenols are significant components of creosote and tar, which have traditionally been used for wood preservation. Although aerobic growth at the expense of aromatic hydrocarbons has been known for many years (Söhngen 1913; Tausson 1927; Gray and Thornton 1928), it was many years later before details of the ring-fission reactions began to emerge. Two converging lines of investigations have examined them in detail: (a) the degradation of the monocyclic aromatic hydrocarbons benzene, toluene, and the xylenes and (b) the degradation of oxygen-substituted compounds such as benzoate, hydroxybenzoates, and phenols. As a result of this activity, the pathways of degradation and their regulation are now known in considerable detail, and ever-increasing attention has been directed to the degradation of polycyclic aromatic hydrocarbons. Since many of these metabolic sequences recur in the degradation of a wide range of aromatic compounds, a brief sketch of the principal reactions may conveniently be presented here. Reviews that include almost all aspects have been given, for example, by Hopper (1978), Cripps and Watkinson (1978), Ribbons and Eaton (1982), Gibson and Subramanian (1984), Smith (1996), and Neilson and Allard (1998). In addition, developments in regulatory aspects have been presented in Rothmel et al. (1991), van der Meer et al. (1992), and Parales and Harwood (1993).

MONOCYCLIC ARENES Aerobic Conditions It is important at the outset to appreciate two important issues. 1. For complete degradation of an aromatic hydrocarbon to occur, it is necessary that the products of ring oxidation and fission can be further degraded to molecules that enter anabolic and energy-producing reactions. 385

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Environmental Degradation and Transformation of Organic Chemicals H

OH OH H

(a) H

(b)

OH H OH

FIGURE 8.1 Alternative pathways for the oxidative metabolism of naphthalene by (a) bacteria and (b) fungi. CO2H

OH

OH

OCH3

OCH3

OH

OH

OH

O

O O

O

OH

CO2H CO2H

CO2H CO2H

OH

FIGURE 8.2

Biodegradation of vanillic acid by fungi.

2. Essentially different mechanisms operate in bacteria and fungi, and these differences have important consequences. In bacteria, the initial reaction is carried out by dioxygenation and results in the synthesis of a cis-1,2-dihydro-1,2-diol, which is then dehydrogenated to a catechol before ring fission mediated by dioxygenases. However, in fungi, the first reaction is monooxygenation to an epoxide followed by hydrolysis to a trans-1,2-dihydro-1,2-diol and rearrangement to a phenol. Ring fission of polycyclic aromatic hydrocarbons does not generally occur in fungi, so that these reactions are essentially biotransformations. These reactions are schematically illustrated in Figure 8.1. 3. However, both fungi and yeasts are able to degrade simpler substituted aromatic compounds such as 3,4-dihydroxybenzoate (Cain et al. 1968), which can be produced as intermediates. The pathways that are used differ from those used by bacteria. Two examples are given, both of which involve concomitant hydroxylation and decarboxylation: (a) the degradation of 3,4dihydroxybenzoate by the yeast Trichosporon cutaneum produces 1,2,4-trihydroxybenzene prior to ring fission (Anderson and Dagley 1980), and (b) the pathway used for vanillate (Figure 8.2) (Ander et al. 1983). The oxidative decarboxylation of a range of 4-hydroxybenzoates to 1,4-dihydroxy compounds has been examined in yeasts, and is carried out by an FAD-dependent monooxygenase in Candida parapsilosis CBS604 (Eppink et al. 1997). Benzene Benzene is one of a group of related aromatic monocyclic hydrocarbons (BTEX—benzene, toluene, ethylbenzene, and xylene), and since these are water soluble, there has been concern for their dissipation and persistence in groundwater under both aerobic and anaerobic conditions. Although aerobic growth at the expense of benzene was established many years ago, the pathway for its degradation was established only much later. The aerobic degradation of benzene by bacteria is

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initiated by formation of a cis-dihydrodiol, followed by dehydrogenation to catechol (Gibson et al. 1968). In a strain of Moraxella sp., subsequent intradiol fission by catechol-1,2-dioxygenase produced muconate that was degraded by the 3-ketoadipate pathway to succinate and acetate (Högn and Jaenicke 1972). In other organisms including Pseudomonas putida, extradiol ring fission by catechol-2,3-dioxygenase produced 2-hydroxymuconate semialdehyde that was degraded to pyruvate and acetaldehyde. Formation of the dihydrodiol in a mutant strain of Pseudomonas putida involves incorporation of both atoms of 18O2 (Gibson et al. 1970). It is mediated by a dioxygenase that has been purified and shown to be a soluble enzyme consisting of three protein components (Axcell and Geary 1975) encoded by the genes bedA, bedB, and bedC. A flavoprotein reductaseBED containing FAD accepts electrons from NADH and transfers them to a [2Fe–2S] Rieske ferredoxinBED, while the terminal oxygenase is a high-molecular-weight protein consisting of two large subunits containing two [2Fe–2S] Rieske ferredoxins (ISPBED) (Crutcher and Geary 1979). The dihydrodiol dehydrogenase that converts the dihydrodiol into catechol is a large enzyme with a molecular weight of 440,000 (Axcell and Geary 1973). In Pseudomonas putida ML2, the genes encoding benzene dioxygenase are carried on a 112-kb plasmid (Tan and Mason 1990). As an alternative, successive monooxygenation of benzene to phenol, catechol, and 1,2,3trihydroxybenzene may be accomplished by the toluene 4-monooxygenase of Pseudomonas mendocina strain JKR1 and the 3-monooxygenase of Ralstonia (Pseudomonas) pickettii strain PKO1 (Tao et al. 2004). Toluene The degradation of toluene has been studied extensively in strains of Pseudomonas putida, and details of the three different pathways have been resolved. 1. In strains containing the TOL plasmid, degradation proceeds by successive oxidation of the methyl group to hydroxymethyl, aldehyde, and carboxylate, followed by dioxygenation of benzoate to catechol and extradiol ring fission (Figure 8.3a) (Keil and Williams 1985). The TOL plasmid also carries the genes for the degradation of 1,3- and 1,4-dimethylbenzenes (Assinder and Williams 1990), and involves comparable oxidations of a single methyl group (Davey and Gibson 1974; Williams and Worsey 1976). These carboxylic acids are then converted into catechols by benzoate dioxygenase and then undergo extradiol ring fission (Williams and Worsey 1976). In Pseudomonas desmolytica, n-propyl benzene is degraded by alternative pathways, both of which terminate in ring fission: (a) oxidation of the side chain to produce benzoate or (b) by dioxygenation and dehydrogenation to 3-n-propyl catechol (Jigami et al. 1979). 2. The genes for toluene degradation may also be located on the chromosome, when a different pathway is followed—dioxygenation by toluene 2,3-dioxygenase to (+)-cis-(2R,3S) 2,3-dihydro-2,3-dihydroxytoluene (Boyd et al. 1992; Ziffer et al. 1977), dehydrogenation to 3-methyl catechol followed by extradiol ring fission (Figure 8.3b). The multicomponent dioxygenase enzymes encoded by the genes todA, todB, todC1, and todC2 are analogous to those involved in benzene degradation: a flavoprotein reductaseTOL that accepts electrons from NADH, and a ferredoxinTOL that transfers them to the terminal dioxygenase (ISPTOL) which consists of two subunits (Zylstra and Gibson 1989). The dihydrodiol dehydrogenase has, however, a much lower molecular weight (104 kDa) than that for the corresponding benzene dihydrodiol dehydrogenase (Rogers and Gibson 1977). 3. Degradation of toluene may also take place in different species of Pseudomonas by initial monooxygenation at the 2-, 3-, or 4-positions. The situation for toluene monooxygenases has been confusing. However, it is now recognized that only the following two initial reactions take place: a. Toluene-2-monooxygenase of Burkholderia cepacia G4 hydroxylates toluene at the ortho position (Shields et al. 1995) and then further to 3-methylcatechol.

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CH2OH

CHO

CO2H

OH

OH OH

(a)

CO2H CHO

CH3

CH3

CH3

H OH

CH3 OH

O OH

(b) OH H

CH3

CH3

OH

CO2H

CO2H

CO2H

CO2H

(c) OH OH

OH

CO2H CO2H

OH

FIGURE 8.3 Aerobic degradation of toluene by (a) side-chain oxidation, (b) dioxygenation of the ring, and (c) monooxygenation. (From Neilson, A.H. and Allard, A.-S. The Handbook of Environmental Chemistry, Springer, 1998. With permission.)

b. The formation of 4-methylphenol by the three-component monoxygenase in P. mendocina strain KR1 plausibly involves an arene oxide intermediate since it takes place with an NIH shift of a proton to the 3-position (Whited and Gibson 1991a,b). Further degradation involves oxidation to 4-hydroxybenzoate, hydroxylation to 3,4dihydroxybenzoate followed by intradiol ring fission (Figure 8.3c). c. Although the monooxygenase from Ralstonia pickettii was originally reported to form 3-hydroxytoluene, it has been shown that 4-hydroxytoluene is produced as the major product (Fishman et al. 2004). Other examples of monooxygenation of arenes include the following: a. The soluble methane monooxygenase from Methylococcus capsulatus produced 4-ethylphenol and 1-phenylethanol from ethylbenzene, and both 1- and 2-hydroxynaphthalene from naphthalene (Dalton et al. 1981); the formation of the 4-ethylphenol is accompanied by an NIH shift suggesting the involvement of an intermediate arene oxide. b. Under conditions of active NH4+ oxidation, cells of Nitrosomonas europaea can transform ethylbenzene to 4-ethylphenol and acetophenone (Keener and Arp 1994). c. In cell extracts, the cytochrome P450 system of Streptomyces griseus transforms benzene to phenol, toluene to 2-methylphenol, biphenyl to 2- and 4-hydroxybiphenyl, styrene to 4-hydroxystyrene, and naphthalene to naphth-1-ol (Sariaslani et al. 1989). Xylenes The pathway for the degradation of the xylenes depends critically on the orientation of the methyl groups, and o-xylene is considered to be the most recalcitrant since xylene monooxygenase cannot hydroxylate one of its methyl groups: a. The degradation of m- and p-xylene in organisms carrying the TOL plasmid is initiated, as for toluene, by oxidation of one of the methyl groups to m- or p-toluic acid. These are then degraded to 3- or 4-methylcatechol by dioxygenation and decarboxylation (Davey and Gibson 1974).

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b. The degradation of o-xylene by chromosomal genes in Pseudomonas stutzeri OX1 takes place, however, by successive hydroxylation to 3,4-dimethylcatechol (Bertoni et al. 1998) that is degraded by dioxygenation and ring fission, with the loss of acetate to propionaldehyde and pyruvate. c. Degradation of o-xylene by Rhodococcus sp. DK17 is initiated by dioxygenation to an unstable o-xylene cis-3,4-dihydrodiol. This may be dehydrogenated to 3,4-dimethylcatechol followed by extradiol ring fission, or undergo dehydration to produce 2,3- and 3,4dimethylphenol (Kim et al. 2004). The degradation of BTEX both individually or in admixture has been shown in a lignindegrading white-rot fungus under nonlignolytic conditions, and was confirmed with ring-labeled toluene (Yadav and Reddy 1993). Interest in fungal transformation of PAHs is noted in the next part of this chapter, and illustrative examples of the hydroxylation of monocyclic arenes include the following (Smith and Rosazza 1983): a. Benzene to phenol by Penicillium chrysogenum, Cunninghamella blakesleeana, C. baineri, and Gliocladium deliquens b. Toluene to 2- and 4-methylphenol by Penicillium chrysogenum, Rhizopus stolonifer, and C. baineri c. Trans-stilbene to 4-hydroxy- and 4,4′-dihydroxystilbene by C. baineri Anaerobic Conditions Concern has arisen that leakage of BTEX from gasoline may reach groundwater where its fate is determined by anaerobic reactions. Although the degradation of benzene can occur anaerobically under sulfate-reducing, denitrifying, methanogenic, and iron (III)-reducing conditions (Mancini et al. 2003), the complete pathway for these reactions has not been established. Details of intermediates have been obtained using enrichment cultures using 13C6-benzene: labeled phenol, benzoate, and toluene were detected and the kinetics of their formation were determined (Ulrich et al. 2005). The pathway for Dechloromonas aromatica strain RCB involves established reactions—hydroxylation to phenol followed by carboxylation and dehydroxylation to benzoate (Chakraborty and Coates 2005). The last two reactions are analogous to those involved in the anaerobic degradation of phenol by Thauera aromatica (Breese and Fuchs 1998; Schühle and Fuchs 2004). Dechloromonas aromatica strain RCB is also able to degrade toluene, ethylbenzene, and xylenes partially (Chakraborty et al. 2005). Perchlorate could also be used as an electron acceptor for the degradation of benzene, while benzene and toluene could be degraded concurrently. The anaerobic mineralization of BTEX in Fe-reducing cultures has been observed (Jahn et al. 2005). The pathways for the degradation of toluene and xylene under denitrifying and sulfatereducing conditions have been studied most extensively, and they take place by reactions quite different from those used by aerobic bacteria. As an example, two anaerobes affiliated with known sulfate-reducing bacteria isolated from enrichments with crude oil were able to grow at the expense of a number of alkylated benzenes—strain oXyS1 with toluene, o-xylene, and o-ethyltoluene and strain mXyS1 with toluene, m-xylene, and m-ethyltoluene (Harms et al. 1999). Several mechanisms have been established for the anaerobic degradation of toluene: 1. Dehydrogenation of the methyl group. Strains of denitrifying bacteria have been shown to degrade toluene in the absence of oxygen using N2O as electron acceptor (Schocher et al. 1991), and the data are consistent with a pathway involving successive oxidation of the ring methyl group with the formation of benzoate. The details of this pathway involving benzyl alcohol and benzaldehyde have been clearly demonstrated with a strain of Thauera (Pseudomonas) sp. under denitrifying conditions (Altenschmidt and Fuchs 1992).

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This pathway is supported by the demonstration of benzyl alcohol dehydrogenase, benzaldehyde dehydrogenase, benzoyl-CoA ligase, and benzoyl-CoA reductase activities in cell extracts (Biegert and Fuchs 1995). The benzyl alcohol dehydrogenase from benzyl alcohol–grown cells was similar in many of its properties to those from the aerobic bacteria Acinetobacter calcoaceticus and Pseudomonas putida (Biegert et al. 1995). 2. Condensation reactions. It has been suggested that the degradation of toluene could proceed by condensation with acetate to form phenylpropionate and benzoate before ring fission (Evans et al. 1992). These are produced as terminal metabolites during anaerobic degradation of toluene by sulfate-reducing enrichment cultures (Beller et al. 1992). A mechanism for the oxidation of the methyl group has been proposed for Azoarcus tolulyticus, and involves a condensation reaction with acetyl-CoA followed by dehydrogenation to cinnamoyl-CoA (Migaud et al. 1996). This is then either transformed into benzylsuccinate and benzylfumarate, which are apparently terminal metabolites that have also been isolated from a denitrifying organism (Evans et al. 1992), or by further degradation to benzoyl-CoA, which is the substrate for the ring reductase (Figure 8.4). The later reactions are also parts of the fumarate pathway. 3. Reactions with fumarate. An alternative sequence involves the direct formation of benzylsuccinate by the reaction of toluene with fumarate in Thauera aromatica (Biegert et al. 1996), and in a strain designated Pseudomonas sp. strain T2 (Beller and Spormann 1997a,b). It has been shown in the denitrifying strain T2 that the reaction between toluene and fumarate is stereospecific, yielding (+)-benzylsuccinate, and that the proton abstracted from toluene is incorporated into the benzylsuccinate (Figure 8.5) (Beller and Spormann 1998). The synthesis of benzylsuccinate during the anaerobic degradation of toluene under denitrifying conditions involves activation of toluene to a benzyl radical by a mechanism involving glycyl radical catalysis (Leuthner et al. 1998; Verfürth et al. 2004). The amino HO2C CH2

CH3

CO2H

CO−SCoA

HO2C + CO2H

FIGURE 8.4

Anaerobic degradation of toluene via benzylsuccinate.

CO2H

COSCoA HO

CH3

CHO

CO2H

CO−SCoA HO

CO−SCoA

CO2H

CO2H

CO−SCoA HO

O

HO CO2H

FIGURE 8.5

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Anaerobic degradation of toluene.

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acid sequence of the large subunit of the purified enzyme from Thauera aromatica showed a high level of homology to glycine radical enzymes and particularly to pyruvate formate lyase. It has been shown (Coschigano et al. 1998) in two mutants of strain T1 that the genes tutD and tutE involved in the anaerobic degradation of toluene encode proteins with molecular masses of 97.6 and 41.3 kDa, which possess homologies to pyruvate formate lyase and its activating enzyme, respectively. A free radical at Gly734 is involved in the operation of pyruvate formate lyase (Wagner et al. 1992) and there is, therefore, a formal similarity in at least one step in the two reaction pathways. The resulting benzylsuccinate is then metabolized to benzoyl-CoA followed by its metabolism, which has already been noted. Comparable pathways are used by Thauera aromatica and Azoarcus sp. for the degradation of m-xylene under denitrifying conditions (Verfürth et al. 2004). A sulfate-reducing bacterium strain PRTOL1 also catalyzes the reaction between toluene and fumarate with the formation of benzylsuccinate, so that this is apparently a widespread reaction (Beller and Spormann 1997a,b). It is also used by Desulfobacula toluolytica (Rabus and Heider 1998), for the degradation of o- and m-xylene by a sulfate-reducing strain of Desulfotomaculum (Morasch et al. 2004), and for xylene and ethylbenzene under sulfate-reducing conditions (Kniemeyer et al. 2003). The analogous succinates and fumarates that would be produced from dimethyl benzenes have been used as markers for the anaerobic degradation of these substrates in an aquifer (Beller et al. 1995; Beller 2002). Strains may be able to regulate the pathway of degradation of toluene to the availability of oxygen: aerobic degradation by dioxygenation or, in the absence of oxygen and under denitrifying conditions, by the benzylsuccinate pathway (Shinoda et al. 2004). 4. Other reactions. a. Under denitrifying conditions that have been extensively examined in strain EbN1, whereas the degradation of toluene proceeds via the fumarate pathway, the degradation of ethylbenzene involves dehydrogenation at the benzylic carbon atom, dehydrogenation, carboxylation, and hydrolysis (Rabus et al. 2002; Kühner et al. 2005) (Figure 8.6). The dehydrogenase is a soluble periplasmic trimeric enzyme that contains molybdenum, iron, acid-labile sulfur, and the molybdopterin cofactor (Johnson et al. 2001). b. Anaerobic hydroxylation of toluene to 4-hydroxytoluene followed by oxidation of the methyl group and dehydroxylation to benzoate has been suggested (Rudolphi et al. 1991). Although the role of these in degradation has not been clearly established, at least two analogies can be suggested: (i) carboxylation and dehydroxylation to 3methylbenzoate that is produced from 1,3-dimethylbenzene by Thauera sp. strain K172 under denitrifying conditions (Biegert and Fuchs 1995) and (ii) oxidation to 4hydroxybenzoate and further degradation. c. An alternative pathway for toluene includes side-chain carboxylation to phenyl acetate and oxidation via phenylglyoxylate to benzoate, which is an established pathway for phenyl acetate itself (Dangel et al. 1991).

O CH3 CH2

FIGURE 8.6

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HO

CH3

CH3 CO

O

C

O

CH2

O

SCoA

Anaerobic degradation of ethylbenzene.

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Styrene This is used in large amounts for the production of polymers, and attention has been directed to the degradation of the volatile monomer that may be discharged into the environment or collected in biofilters. The bacterial degradation and transformation of styrene has attracted considerable attention (Warhurst and Fewson 1994), and several pathways have been described for bacteria: 1. In several bacteria, ring dioxygenation with the production of the cis-2,3-dihydrodiol takes place, leaving the vinyl group intact. In Rhodococcus rhodochrous strain NCIMB 13259, 2-vinyl-cis,cis-muconate is then produced by catechol 1,2-dioxygenase as a terminal metabolite, although complete degradation is possible owing to the existence of catechol 2,3-dioxygenase activity that is present (Warhurst et al. 1994). 2. Naphthalene dioxygenase from Pseudomonas sp. strain NCIB 9816-4 produces (R)-1phenyl-1,2-ethandiol by stereospecific dioxygenation of the vinyl group (Lee and Gibson 1996). In a formally analogous reaction, the 4-methoxybenzoate monooxygenase from Pseudomonas putida strain DSM 1868 brings about dioxygenation of 4-vinyl benzoate to 4-(1,2-dihydroxyethyl)benzoate (Wende et al. 1989). 3. Monooxygenation of the vinyl group with formation of the epoxide takes place with a number of bacteria including Methylococcus capsulatus strain Bath (Colby et al. 1977) and Nitrosomonas europaea (Keener and Arp 1994, p. 202). The degradation of styrene oxide by Xanthobacter sp. strain 124X produced phenylacetaldehyde by the activity of an isomerase (Hartmans et al. 1989), and in an unclassified Gram-positive organism—though not in the 124X strain—the degradation of styrene is accomplished by an FAD-dependent monooxygenase (Hartmans et al. 1990). Further metabolism of phenylacetaldehyde in Corynebacterium sp. strain ST-10 is carried out by an aldehyde reductase that is able to accept a wide range of aldehyde substrates (Itoh et al. 1997). In Pseudomonas sp. strain Y2, phenylacetaldehyde enters the lower pathway for degradation of phenylacetate via the coenzyme A-ester (del Peso-Santos et al. 2006). 4. The initial steps in the degradation of 2′-methyl-4-methoxystyrene (trans-anethole) by a strain of Arthrobacter aurescens are comparable to those already noted, except that the epoxide is hydrolyzed to a diol to produce 4-methoxybenzoate (Shimoni et al. 2002). It is appropriate here to note that styrene is transformed by the black yeast Exophilia jeanselmei to phenylacetate by a pathway similar to that of the Xanthobacter sp. already noted. The initial monooxygenation was carried out by a cytochrome P450, and phenylacetate was further metabolized to 2-hydroxy- and 2,5-dihydroxyphenylacetate (Cox et al. 1996). Stilbene Substituted stilbenes are formed during the production of wood pulp, and are components of fluorescent whiting agents. The latter are considered recalcitrant although there are a few examples of their biodegradation. The degradation of stilbenes by oxidative fission of the Ar–C=C–Ar bond has been described for a lignin model compound (Habu et al. 1989b) putatively by dioxygenation (Habu et al. 1989a), and presumptively for α-methyl-4,4′-dihydroxy stilbene that is an intermediate in the degradation of bisphenol-A (Spivack et al. 1994). Synthetic Applications It has already been shown that biodegradation of many aromatic compounds proceeds by initial dioxygenation to cis-dihydrodiols, followed by dehydrogenation and ring fission. The high enantiomeric purity of cis-dihydrodiols produced by bacterial dioxygenases has been emphasized. Mutant strains that lack dehydrogenase activity produce only dihydrodiols, and there has been increasing interest in developing the use of these as synthons for the production of novel compounds that would not be readily available by conventional chemical synthesis.

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For dihydrodiols derived from substituted benzenes, the key to their significance lies in the availability of two adjacent chiral centers with an established absolute stereochemistry. The dihydrodiol from benzene is, of course, the meso compound, although enantiomers produced by subsequent reaction with a chiral reagent are readily separated. There are useful reviews containing numerous applications (Carless 1992; Ribbons et al. 1989), many of which involve, in addition, the use of cis-fluoro-, cis-chloro-, or cis-bromobenzene-2,3-dihydrodiols. Only a few illustrative syntheses using benzene and toluene cis-dihydrodiols are given below: 1. Although the product from the transformation of toluene by mutants of Pseudomonas putida lacking dehydrogenase activity is the cis-2R,3S dihydrodiol, the cis-2S,3R dihydrodiol has been synthesized from 4-iodotoluene by a combination of microbiological and chemical reactions. P. putida strain UV4 was used to prepare both enantiomers of the cis-dihydrodiol, and iodine was chemically removed using H2 -Pd/C. Incubation of the mixture of enantiomers with P. putida NCIMB 8859 selectively degraded the 2R,3S compound to produce toluene cis-2S,3R dihydrodiol (Allen et al. 1995). 2. Racemic pinitol from benzene cis-dihydrodiol benzoate by successive epoxidation and osmylation (Figure 8.7a) (Ley et al. 1987). 3. Conduramine A1 tetraacetate using an activated nitroso-mannose derivative (Figure 8.7b) (Werbitzky et al. 1990). 4. (–)-Laminitol from toluene cis-dihydrodiol by successive epoxidations (Figure 8.7c) (Carless and Oak 1991). 5. Analogs of conduritols from toluene cis-dihydrodiol by reaction with singlet oxygen followed by scission with thiourea (Figure 8.7d) (Carless et al. 1989). OH

O

OBz

OBz

MeO

OH OBz

MeO

OH

OBz

HO

OH

(a) OBz

OBz NH

NHAc

O OAc

OH OAc

+

(b) OAc

OAc

Cl

OAc OAC

OAc OMe O

O

Me

HO O

Me

HO

HO

Me

OH

HO

OH

OH

HO

OH

(c) O

O

OH Me

Me OH O (d) OH

O

HO

Me OH

OH OH

OH OH

FIGURE 8.7 Examples of chemical syntheses based on cyclohexadiene cis dihydrodiols: (a) pinitol, (b) conduramine A, (c) (–)-laminitol, and (d) conduritol analogs. (From Neilson, A.H. and Allard, A.-S. The Handbook of Environmental Chemistry, Springer, 1998. With permission.)

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There is clearly enormous potential using other cis dihydrodiols produced from benzocycloalkenes, or from the numerous dihydrodiols produced from polycyclic carbocyclic and heterocyclic substrates.

REFERENCES Allen CCR, DR Boyd, H Dalton, ND Sharma, I Brannigan, NA Kerley, GN Sheldrake, SC Taylor (1995) Enantioselective bacterial biotransformation routes to cis-diol metabolites of monosubstituted benzenes, naphthalene and benzocycloalkenes of either absolute configuration. J Chem Soc Chem Commun 117–118. Altenschmidt U, G Fuchs (1992) Anaerobic toluene oxidation to benzylalcohol and benzaldehyde in a denitrifying Pseudomonas sp. J Bacteriol 174: 4860–4862. Ander P, K-E Eriksson, H-S Yu (1983) Vanillic acid metabolism by Sporotrichium pulverulentum: evidence for demethoxylation before ring-cleavage. Arch Microbiol 136: 1–6. Anderson JJ, S Dagley (1980) Catabolism of aromatic acids in Trichosporon cutaneum. J Bacteriol 141: 534–543. Assinder SJ, PA Williams (1990) The TOL plasmids: determinants of the catabolism of toluene and the xylenes. Adv Microb Physiol 31: 1–69. Axcell BC, PJ Geary (1973) The metabolism of benzene by bacteria. Purification and some properties of the ezyme cis-1,2-dihydroxycyclohexa-3,5-diene (nicotinamide adenine dinucleotide) oxidoreductase (cis-benzene glycol dehydrogenase). Biochem J 136: 927–934. Axcell BC, PJ Geary (1975) Purification and some properties of a soluble benzene-oxidizing system from a strain of Pseudomonas. Biochem J 146: 173–183. Beller HR (2002) Analysis of benzylsuccinates in groundwater by liquid chromatography/tandem mass spectrometry and its use for monitoring in situ BTEX biodegradation. Environ Sci Technol 36: 2724–2728. Beller HR, AM Spormann (1997a) Anaerobic activation of toluene and o-xylene by addition to fumarate in denitrifying strain T. J Bacteriol 179: 670–676. Beller HR, AM Spormann (1997b) Benzylsuccinate formation as a means of anaerobic toluene activation by sulfate-reducing strain PRTOL1. Appl Environ Microbiol 63: 3729–3731. Beller HR, AM Spormann (1998) Analysis of the novel benzylsuccinate synthase reaction for anaerobic toluene activation based on structural studies of the product. J Bacteriol 180: 5454–5457. Beller HR, M Reinhard, D Grbic-Galic (1992) Metabolic by-products of anaerobic toluene degradation by sulfate-reducing enrichment cultures. Appl Environ Microbiol 58: 3192–3195. Beller HR, W-H Ding, M Reinhard (1995) Byproducts of anaerobic alkylbenzene metabolism useful as indicators of in situ bioremediation. Environ Sci Technol 29: 2864–2870. Bertoni G, M Martino, E Galli, P Barbieri (1998) Analysis of the gene cluster encoding toluene/o-xylene monooxygenase from Pseudomonas stutzeri OX1. Appl Environ Microbiol 64: 3626–3632. Biegert T, G Fuchs (1995) Anaerobic oxidation of toluene (analogues) to benzoate (analogues) by whole cells and by cell extracts of a denitrifying Thauera sp. Arch Microbiol 163: 407–417. Biegert T, G Fuchs, J Heider (1996) Evidence that anaerobic oxidation of toluene in the denitrifying bacterium Thauera aromatica is initiated by formation of benzylsuccinate from toluene and fumarate. Eur J Biochem 238: 661–668. Biegert T, U Altenschmidt, C Eckerskorn, G Fuchs (1995) Purification and properties of benzyl alcohol dehydrogenase from a denitrifying Thauera sp. Arch Microbiol 163: 418–423. 1 Boyd DR, ND Sharma, R Boyle, RAS McMordie, J Chima, H Dalton (1992) A H NMR method for the determination of enantiomeric excess and absolute configuration of cis-dihydrodiol metabolites of polycyclic arenes and heteroarenes. Tetrahedron Lett 33: 1241–1244. Breese K, M Boll, J Alt-Mörbe, H Schäggrer, G Fuchs (1998) Genes encoding the benzoyl-CoA pathways of anaerobic aromatic metabolism in the bacterium Thauera aromatica. Eur J Biochem 256: 148–154. Cain RB, RF Bilton, JA Darrah (1968) The metabolism of aromatic acids by micro-organisms. Metabolic pathways in the fungi. Biochem J 108: 797–828. Carless HAJ (1992) The use of cyclohexa-3,5-diene-1,2-diols in enantiospecific synthesis. Tetrahedron Asymmetry 3: 795–826. Carless HAJ, Billinge JR, Oak OZ (1989) Photochemical routes from arenes to inositol intermediates: the photo-oxidation of substituted cis-cyclohexane-3,5-diene-1,2-diols. Tetrahedron Lett 30: 3113–3116.

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Carless HAJ, Oak OZ (1991) Total synthesis of (–)-laminitol (1D-4C-methyl-myo-inositol) via microbial oxidation of toluene. Tetrahedron Lett 32: 1671–1674. Chakraborty R, JD Coates (2005) Hydroxylation and carboxylation—two crucial steps of anaerobic benzene degradation by Dechloromonas strain RCB. Appl Environ Microbiol 71: 5427–5432. Chakraborty R, SM O’Connor, E Chan, JD Coates (2005) Anaerobic degradation of benzene, toluene, ethylbenzene, and xylene compounds by Dechloromonas strain RCB. Appl Environ Microbiol 71: 8649–8655. Colby J, DI Stirling, H Dalton (1977) The soluble methane mono-oxygenase of Methylococcus capsulatus (Bath). Its ability to oxygenate n-alkanes, n-alkenes, ethers, and alicyclic, aromatic, and heterocyclic compounds. Biochem J 165: 395–402. Coschigano PW, TS Wehrman, LY Young (1998) Identification and analysis of genes involved in anaerobic toluene metabolism by strain T1: putative role of a glycine free radical. Appl Environ Microbiol 64: 1650–1656. Cox HHJ, BW Faber, VNM van Heiningen, H Radhoe, HJ Doddema, W Harder (1996) Styrene metabolism in Exophilia jeanselmei and involvement of a cytochrome P-450-dependent styrene monooxygenase. Appl Environ Microbiol 62: 1471–1474. Cripps RE, RJ Watkinson (1978) Polycyclic hydrocarbons: metabolism and environmental aspects. In Developments in biodegradation of hydrocarbons-1 (Ed RJ Watkinson), pp. 113–134. Applied Science Publishers Ltd, London. Crutcher SE, PJ Geary (1979) Properties of the iron-sulphur proteins of the benzene dioxygenase system from Pseudomonas putida. Biochem J 177: 393–400. Dalton H, BT Gording, BW Watyers, R Higgins, JA Taylor (1981) Oxidations of cyclopropane, methylcyclopropane, and arenes with the monooxygenase system from Methylococccus capsulatus. J Chem Soc Chem Commun 482–483. Dangel W, R Brackmann, A Lack, M Mohamed, J Koch, J Oswald, B Seyfried, A Tschech, G Fuchs (1991) Differential expression of enzyme activities initiating anoxic metabolism of various aromatic compounds via benzoyl-CoA. Arch Microbiol 155: 256–262. Davey JF, DT Gibson (1974) Bacterial metabolism of para- and meta-xylene: oxidation of a methyl substituent. J Bacteriol 119: 923–929. del Peso-Santos T, D Bartolomé-Martín, C Fernández, S Alonso, JL Gardía, E Díaz, V Shingler, J Perera (2006) Coregulation by phenylacetyl-coenzyme A-responsive PaaX integrates control of the upper and lower pathways for catabolism of styrene by Pseudomonas sp. strain Y2. J Bacteriol 188: 4812–4821. Eppink MHM, SA Boeren, J Vervoort, WJH van Berkel (1997) Purification and properties of 4-hydroxybenzoate 1-hydroxylase (decarboxylating), a novel flavin adenine dinucleotide-dependent monooxygenase from Candida parapilosis CBS604. J Bacteriol 179: 6680–6687. Evans PJ, W Ling, B Goldschmidt, ER Ritter, LY Young (1992) Metabolites formed during anaerobic transformation of toluene and o-xylene and their proposed relationship to the initial steps of toluene mineralization. Appl Environ Microbiol 58: 496–501. Fishman A, Y Tao, TK Wood (2004) Toluene 3-monooxygenase of Ralstonia pickettii PKO1 is a parahydrxylating enzyme. J Bacteriol 186: 3117–3123. Gibson DT, GE Cardini, FC Masales, RE Kallio (1970) Incorporation of oxygen-18 into benzene by Pseudomonas putida. Biochemistry 9: 1631–1635. Gibson DT, JR Koch, RE Kallio (1968) Oxidative degradation of aromatic hydrocarbons. I. Enzymatic formation of catechol from benzene. Biochemistry 9: 2653–2662. Gibson DT, V Subramanian (1984) Microbial degradation of aromatic hydrocarbons. In Microbial degradation of organic compounds (Ed DT Gibson), pp. 181–252. Marcel Dekker Inc, New York. Gray PHH, G Thornton (1928) Soil bacteria that decompose certain aromatic compounds. Centralbl Bakteriol Parasitenkd Infektionskr (2 Abt) 73: 74–96. Habu N, M Samejima, T Yoshimoto (1989a) A novel dioxygenase responsible for the Cα-Cβ cleavage of lignin model compounds from Pseudomonas sp. TMY1009. Mokuzai Gakkaishi 35: 26–29. Habu N, M Samejima, T Yoshimoto (1989b) Metabolism of a diarylpropane type lignin model by Pseudomonas sp. TMY 1009. Mokuzai Gakkaishi 35: 348–355. Harms G, K Zengle, R Rabus, F Aeckersberg, D Minz, R Rosselló-Mora, F Widdel (1999) Anaerobic oxidation of o-xylene, m-xylene, and homologous alkylbenzenes by new types of sulfate-reducing bacteria. Appl Environ Microbiol 65: 999–1004. Hartmans S, JP Smits, MJ van der Werf, F Volkering, JAM de Bont (1989) Metabolism of styrene oxide and 2-phenylethanol in the styrene-degrading Xanthobacter strain 124X. Appl Environ Microbiol 55: 2850–2855.

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Hartmans S, MJ van der Werf, JAM de Bont (1990) Bacterial degradation of styrene involving a novel flavin adenenine dinucleotide-dependent styrene monooxygenase. Appl Environ Microbiol 41: 1045–1054. Harwood CS, G Burchardt, H Herrmann, G Fuchs (1999) Anaerobic metabolism of aromatic compounds via the benzoyl-CoA pathway. FEMS Microbiol Rev 22: 439–458. Högn T, L Jaenicke (1972) Benzene metabolism of Moraxella species. Eur J Biochem 30: 369–375. Hopper DJ (1978) Microbial degradation of aromatic hydrocarbons. In Developments in biodegradation of hydrocarbons-1 (Ed RJ Watkinson), pp. 85–112. Applied Science Publishers Ltd, London. Itoh N, R Morihama, J Wang, K Okada, N Mizuguchi (1997) Purification and characterization of phenylacetaldehyde reductase from a styrene-assimilating Corynebacterium strain, ST10. Appl Environ Microbiol 63: 3783–3788. Jahn MK, SB Haderlin, RU Meckenstock (2005) Anaerobic degradation of benzene, toluene, ethylbenzene, and o-xylene in sediment-free iron-reducing enrichment cultures. Appl Environ Microbiol 71: 3355–3358. Jigami Y, Y Kawasaki, T Omori, Y Minoda (1979) Coexistence of different pathways in the metabolism of n-propylbenzene by Pseudomonas sp. Appl Environ Microbiol 38: 783–788. Johnson HA, DA Pelletier, AM Spormann (2001) Isolation and characterization of anaerobic ethylbenzene dehydrogenase, a novel Mo-Fe-S enzyme. J Bacteriol 183: 4536–4542. Keener WK, DJ Arp (1994) Transformations of aromatic compounds by Nitrosomonas europaea. Appl Environ Microbiol 60: 1914–1932. Keil H, PA Williams (1985) A new class of TOL plasmid deletion mutants in Pseudomonas putida MT15 and their reversion by tandem gene amplification. J Gen Microbiol 131: 1023–1033. Kim D, J-C Chae, GJ Zylstra, Y-S Kim, MH Nam, YM Kim, E Kim (2004) Identification of a novel dioxygenase involved in metabolism of o-xylene, toluene, and ethylbenzene by Rhodococcus sp. strain DK17. Appl Environ Microbiol 70: 7086–7092. Kniemeyer O, T Fischer, H Wilkes, FO Glöckner, F Widdel (2003) Anaerobic degradation of ethylbenzene by a new type of marine sulfate-reducing bacterium. Appl Environ Microbiol 69: 760–768. Kühner S, L Wöhlbrand, I Fritz, W Wruck, C Hultschig, P Hufnagel, M Kube, R Reinhardt, R Rabus (2005) Substrate-dependent regulation of anaerobic degradation pathways for toluene and ethylbenzene in a denitrifying bacterium strain EbN1. J Bacteriol 187: 1493–1503. Lee K, DT Gibson (1996a) Stereospecific dihydroxylation of the styrene vinyl group by purified naphthalene dioxygenase from Pseudomonas sp. strain NCIB 9816-4. J Bacteriol 178: 3353–3356. Leuthner B, C Leutwein, H Schulz, P Hörth, W Haehnel, E Schiltz, H Schägger, J Heider (1998) Biochemical and genetic characterisation of benzylsuccinate synthase from Thauera aromatica: a new glycyl-radical catalysing the first step in anaerobic toluene degradation. Mol Microbiol 28: 515–628. Ley SV, F Sternfield, S Taylor (1987) Microbial oxidation in synthesis: a six step preparation of (+/–)-pinitol from benzene. Tetrahedron Lett 28: 225–226. Mancini SA, AC Ulrich, G Lacrampe-Couloume, B Sleep, EA Edwards, BS Lollar (2003) Carbon and hydrogen isotopic fractionation during anaerobic biodegradation of benzene. Appl Environ Microbiol 69: 191–198. Migaud ME, JC Chee-Sandford, JM Tiedje, JW Frost (1996) Benzylfumaric, benzylmaleic, and Z- and E-phenylitaconic acids: synthesis, characterization, and correlation with a metabolite generated by Azoarcus tolulyticus Tol-4 during anaerobic toluene degradation. Appl Environ Microbiol 62: 974–978. Morasch B, B Schink, CC Tebbe, RU Meckenstock (2004) Degradation of o-xylene and m-xylene by a novel sulfate-reducer belonging to the genus Desulfotomaculum. Arch Microbiol 181: 407–417. Neilson AH, A-S Allard (1998) Microbial metabolism of PAHs and heteroarenes. Handb Environ Chem 3J: 1–80. Parales RE, CS Harwood (1993) Regulation of the pcaIJ genes for aromatic acid degradation in Pseudomonas putida. J Bacteriol 175: 5829–5838. Rabus R, J Heider (1998) Initial reactions of anaerobic metabolism of alkylbenzenes in denitrifying and sulfate-reducing bacteria. Arch Microbiol 170: 377–384. Rabus R, M Kube, A Beck, F Widdel, R Reinhardt (2002) Genes involved in the anaerobic degradation of ethylbenzene in a denitrifying bacterium, strain EbN1. Arch Microbiol 178: 506–516. Ribbons DW, RW Eaton (1982) Chemical transformations of aromatic hydrocarbons that support the growth of microorganisms. In Biodegradation and detoxification of environmental pollutants (Ed AM Chakrabarty), pp. 59–84. CRC Press, Boca Raton.

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Ribbons DW, SJC Taylor, CT Evans, SD Thomas, JT Rossiter, DA Widdowson, DJ Williams (1989) Biodegradations yield novel intermediates for chemical synthesis. In Biotechnology and Biodegradation (Eds D Kamely, A Chakrabarty, GS Omenn), pp. 213–245. Gulf Publishing Company, Houston, Texas. Rogers JE, DT Gibson (1977) Purification and properties of cis-toluene dihydrodiol dehydrogenase from Pseudomonas putida. J Bacteriol 130: 1117–1124. Rothmel, RK, DL Shinbarger, MR Parsek, TL Aldrich, AM Chakrabarty (1991) Functional analysis of the Pseudomonas putida regulatory protein CatR: transcriptional studies and determination of the CatR DNA-binding site by hydroxyl-radical footprinting. J Bacteriol 173: 4717–4724. Rudolphi A, A Tschech, G Fuchs G (1991) Anaerobic degradation of cresols by denitrifying bacteria. Arch Microbiol 155: 238–248. Sariaslani FS, MK Trower, SE Buchloz (1989) Xenobiotic transformations by Streptomyces griseus. Dev Ind Microbiol 30: 161–171. Schocher RJ, B Seyfried, F Vazquez, J Zeyer (1991) Anaerobic degradation of toluene by pure cultures of denitrifying bacteria. Arch Microbiol 157: 7–12. Schühle K, G Fuchs (2004) Phenylphosphate carboxylase: a new C–C lyase involved in anaerobic phenol metabolism in Thauera aromatica. J Bacteriol 186: 4556–4567. Shields MS, MJ Reagin, RR Gerger, R Campbell, C Somerville (1995) TOM, a new aromatic degradative plasmid from Burkholderia (Pseudomonas) cepacia G4. Appl Environ Microbiol 61: 1352–1356. Shimoni E, T Baasov, U Ravid, Y Shoham (2002) The trans-anethole degradation pathway in an Arthrobacter sp. J Biol Chem 277: 11866–11872. Shinoda Y, Y Sakai, H Uenishi, Y Uchihashi, A Hiraishi, H Yukawa, H Yurimoto, N Kato (2004) Aerobic and anaerobic toluene degradation by a newly isolated denitrifying bacterium Thauera sp. strain DNT-1. Appl Environ Microbiol 70: 1385–1392. Smith MR (1994) The physiology of aromatic hydrocarbon degrading bacteria. In Biochemistry of Microbial Degradation (Ed C Ratledge), pp. 347–378. Kluwer Academic Publishers, Dordrecht, The Netherlands. Smith RV, JP Rosazza (1983) Microbial models of mammalian metabolism. J Nat Prod 46: 79–91. Söhngen NL (1913) Benzin, Petroleum, Paraffinöl und Paraffin als Kohlenstoff- und Energiequelle für Mikroben. Centralbl Bakteriol Parasitenkd Infektionskr (2 Abt) 37: 595–609. Spivack J, TK Leib, JH Lobos (1994) Novel pathway for bacterial metabolism of bisphenol A. Rearrangements and stilbene cleavage in bisphenol A metabolism. J Biol Chem 269: 7323–7329. Tan H-M, JR Mason (1990) Cloning and expression of the plasmid-encoded benzene dioxygenase from Pseudomonas putida ML2. FEMS Microbiol Lett 72: 259–264. Tao Y, A Fishman, WE Bentley, TK Wood (2004) Oxidation of benzene to phenol, catechol, and 1,2,3trihydroxybenzene by toluene 4-monooxygenase of Pseudomonas mendocina KR 1 and toluene 3-monooxygenase of Ralstonia pickettii PKO1. Appl Environ Microbiol 70: 3814–3820. Tausson WO (1927) Naphthalin als Kohlenstoffquelle für Bakterien. Planta 4: 214–256. Ulrich AC, HR Beller, EA Edwards (2005) Metabolites detected during biodegradation of 13C6 -benzene in nitrate-reducing and methanogenic enrichment cultures. Environ Sci Technol 39: 6681–6691. van der Meer, WM de Vos, S Harayama, AJB Zehnder (1992) Molecular mechanisms of genetic adaptation to xenobiotic compounds. Microbiol Revs 56: 677–694. Verfürth K, AJ Pierik, C Leutwein, S Zorn, J Heider (2004) Substrate specificities and electron paramagnetic resonance properties of benzylsuccinate synthesis in anaerobic toluene and m-xylene metabolism. Arch Microbiol 181: 155–162. Wagner AFV, M Frey, FA Neugebauer, W Schäfer, J Knappe (1992) The free radical in pyruvate formate-lyase is located on glycine-734. Proc Natl Acad USA 89: 996–1000. Warhurst AM, CA Fewson (1994) Microbial metabolism and biotransformations of styrene. J Appl Bacteriol 77: 597–606. Warhurst AM, KE Clarke, RA Hill, RA Holt, CA Fewson (1994) Metabolism of styrene by Rhodococcus rhodochrous NCIMB 13259. Appl Environ Microbiol 60: 1137–1145. Wende P, F-H Bernhardt, K Pfleger (1989) Substrate-modulated reactions of putidamonooxin. The nature of the active oxygen species formed and its reaction mechanism. Eur J Biochem 181: 189–197. Werbitzky O, K Klier, H Felber H (1990) Asymmetric induction of four chiral cenbters by hetero Diels–Alder reaction of a chiral nitrosodienmophile. Liebigs Ann Chem 267–270. Whited GM, DT Gibson (1991a) Separation and partial characterization of the enzymes of the toluene-4monooxygenase catabolic pathway in Pseudomonas mendocina KR1. J Bacteriol 173: 3017–3020.

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Whited GM, DT Gibson (1991b) Toluene-4-monooxygenase, a three-component enzyme system that catalyzes the oxidation of toluene to p-cresol in Pseudomonas mendocina KR1. J Bacteriol 173: 3010–3016. Williams PA, MJ Worsey (1976) Ubiquity of plasmids in coding for toluene and xylene metabolism in soil bacteria: evidence for the existence of new TOL plasmids. J Bacteriol 125: 818–828. Yadav JS, CA Reddy (1993) Degradation of benzene, toluene, ethylbenzene, and xylenes (BTEX) by the lignin-degrading basidiomycete Phanerochaete chrysosporium. Appl Environ Microbiol 59: 756–762. Ziffer H, K Kabuta, DT Gibson, VM Kobal, DM Jerina (1977) The absolute stereochemistry of several cis dihydrodiols microbially produced from substituted benzenes. Tetrahedron 33: 2491–2496. Zylstra GJ, DT Gibson (1989) Toluene degradation by Pseudomonas putida F1. Nucleotide sequence of the todC1C2BADE genes and their expression in Escherichia coli. J Biol Chem 264: 14940–14946.

PART 2 POLYCYCLIC AROMATIC HYDROCARBONS (PAHS) INTRODUCTION Aromatic hydrocarbons are components of petroleum and are produced during the gasification of coal, while naturally occurring partially aromatized steroids and terpenoids which occur in sediments and coal have been used as biomarkers (Simoneit 1998). Concern with polycyclic aromatic hydrocarbons has centered on those that display carcinogenicity to mammals after metabolic activation. Since anoxic sediments may be a sink for PAHs, degradation and biodegradation have been examined under both aerobic and anaerobic conditions. Biochemical aspects of the biodegradation of PAHs have been given in a classic review in Gibson and Subramanian (1984), Smith (1994), and Kanaly and Harayama (2000), while specific aspects of sphingomonads have been presented in Pinyakong et al. (2003). The biodegradation of benzo[a]pyrene by bacteria, fungi, and algae has been addressed in a review within the context of remediation (Juhasz and Naidu 2000). Attention is directed to marine organisms that are able to degrade PAHs, and which generally belong to the genus Cycloclasticus (Geiselbrecht et al. 1998; Kasai et al. 2003; Wang et al. 1996). The aerobic degradation of naphthalene was described many years ago (Tausson 1927; Gray and Thornton 1928). As for the degradation of benzene, it was, however, many years later that the details of the reaction involved were elucidated. Since then, the degradation of naphthalene, anthracene, phenanthrene, fluoranthene, and pyrene has been described in a number of organisms, that of benzo[a]pyrene, benz[a]anthracene, and dibenz[a]anthracene less often, and that of chrysene (Demanèche et al. 2004) and coronene (Juhasz et al. 1997) only occasionally. It should be realized that growth at the expense of PAHs may require only a contribution from the products of partial degradation. Care should therefore be exercised in the interpretation of experiments using partially labeled substrates (Ye et al. 1996). Aerobic degradation is generally initiated by dioxygenation followed by dehydrogenation to catechols that undergo ring fission. For polycyclic arenes, repetition of these reactions is then used for the successive degradation of the other rings, for example, o-phthalate is produced from anthracene (van Herwijnen et al. 2003), phenanthrene (Moody et al. 2001), and pyrene (Vila et al. 2001; Krivobok et al. 2003). Exceptionally, elimination rather than dehydrogenation takes place. For example, the dioxygenation of 7,4′-dihydroxyisoflavone by biphenyl-2,3-dioxygenase from Burkholderia sp. strain LB400 is followed by elimination to the phenol 7,2′,4′trihydroxyisoflavone rather than by dehydrogenation (Seeger et al. 2003). Substantial attention has been directed to details of the degradation of naphthalene on account of interest in the synthesis of chiral compounds using naphthalene-1,2-dioxygenase, and of biphenyl, since biphenyl-2,3-dioxygenase induced by growth with biphenyl has a broad substrate range including congeners of PCBs. Genes for the degradation of naphthalene may be carried on plasmids, for example, in pseudomonads (Dunn and Gunsalus 1973; Dunn et al. 1980; Yen et al. 1988), Novosphingobium aromaticivorans (Romine et al. 1999), and some species of Rhodococcus (Kulakov et al. 2005).

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Aerobic Reactions Carried Out by Bacteria Naphthalene The aerobic degradation of naphthalene and its derivatives has been extensively examined, so that the pathway, biochemistry, and genetics are well established. Naphthalene is readily degraded by many bacteria including Gram-negative pseudomonads, Gram-positive rhodococci, and marine bacteria belonging to the genus Cycloclasticus. Both the details of the initial steps and their enzymology have been elucidated. Genes for the degradation of naphthalene and the intermediate metabolite salicylate are generally carried on plasmids in pseudomonads (Dunn and Gunsalus 1973; Dunn et al. 1980; Yen et al. 1988), and in some species of Rhodococcus (Kulakov et al. 2005). The overall pathway of degradation is shown in Figure 8.8, and is carried out by a sequence of enzymes. The enzymes for the complete sequence of enzymes involved in the oxidative degradation of naphthalene to catechol and for the extradiol (meta) fission pathway for degradation of the catechol are inducible by growth with salicylate in Pseudomonas putida (Austen and Dunn 1980). The genes for degradation in Pseudomonas putida G7 are organized in two operons, the upper encoding those for the conversion of naphthalene into salicylate, and the lower those for the degradation of salicylate to pyruvate and acetate. The nah genes encoding the enzymes on the plasmid NAH7 for conversion of naphthalene into salicylate have been cloned from Pseudomonas putida G1064: nahA, nahB, nahC, nahD, nahE, and nahF for naphthalene dioxygenase, naphthalene cis-dihydrodiol dehydrogenase, 1,2-dihydroxynaphthalene dioxygenase, 2-hydroxychromene-2-carboxylate isomerase, and salicylaldehyde dehydrogenase, respectively (Yen and Serdar 1988). The initial product of naphthalene oxidation in Pseudomonas putida by naphthalene dioxygenase is naphthalene (+)-cis-(1R,2S)-1,2-dihydro-1, 2-diol (Jeffrey et al. 1975). Naphthalene dioxygenase (Patel and Barnsley 1980; Ensley and Gibson 1983), cis-naphthalene dihydrodiol dehydrogenase (Patel and Gibson 1974), and 1,2-dihydroxynaphthalene dioxygenase (Kuhm et al. 1991) have been purified, while further details of the subsequent steps have been added (Eaton and Chapman 1992). The degradation of naphthalene has also been studied in strains of Rhodococcus sp., and two groups of strains have been examined. Differences from pseudomonads have been observed in strain NCIMB 12038, in which the genes are carried on a plasmid: (a) naphthalene is the sole inducer of the whole pathway and (b) the degradation of salicylate proceeds from the activity of salicylate 5-hydroxylase to 2,5-dihydroxybenzoate (gentisate) (Grund et al. 1992; Larkin et al. 1999). In other strains including strain P200, the genes are chromosomal (Kulakov et al. 2005).

H

OH

OH OH

OH H

OH

OH CO2H

OH

OH

CO2H

OH

OH HO

CO2H

FIGURE 8.8 Degradation of naphthalene. (From Neilson, A.H. and Allard, A.-S. The Handbook of Environmental Chemistry, Springer, 1998. With permission.)

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Further details merit brief comment: a. In some strains of pseudomonads, the degradation of the intermediate catechol produced by the activity of salicylate hydroxylase may proceed either by the extradiol or the alternative intradiol fission pathway (Barnsley 1976). b. In Ralstonia (Pseudomonas) sp. strain U2 (Fuenmayor et al. 1998; Zhou et al. 2002), the alternative gentisate pathway for the degradation of the intermediate salicylate is used. c. In a strain of Mycobacterium sp., both dioxygenation to cis-naphthalene-1,2-dihydrodiol and monooxygenation and hydrolysis to trans-naphthalene-1,2-dihydrodiol were encountered, with the latter dominating 25-fold (Kelley et al. 1990). d. In the context of biotechnology, naphthalene dioxygenase from a strain of Pseudomonas sp. also carried out enantiomeric monooxygenation of indan and dehydrogenation of indene (Gibson et al. 1995), and the stereospecific hydroxylation of (R)-1-indanol, (S)-1-indanol to cis-indan-1,3-diol, and trans (1S,3S)-indane-1,3-diol (Lee et al. 1997); the indantriols are also formed by further reactions. Essentially comparable reactions have been observed with Rhodococcus sp. strain NCIMB 12038 (Allen et al. 1997). e. Sphingomonas aromaticivorans carries catabolic genes for the degradation of aromatic hydrocarbons on a plasmid. It is able to degrade not only naphthalene and biphenyl by dioxygenation, but also initiate the degradation of xylenes by monooxygenation of the methyl groups, followed by dioxygenation of the resulting benzoates (Romine et al. 1999). Naphthalene dioxygenase from Pseudomonas putida consists of three proteins: the reductaseNAP (Haigler and Gibson 1990b), a ferredoxinNAP (Haigler and Gibson 1990a), and the two-subunit terminal oxygenase ISPNAP (Ensley and Gibson 1983). They are analogous to those involved in the degradation of benzene and toluene, and are encoded by the genes nadAa, nadAb, and nadAc/nadAd (Simon et al. 1993). However, the reductaseNAD component of naphthalene dioxygenase contains both FAD and a [2Fe–2S] ferredoxin (Haigler and Gibson 1990a), so that naphthalene is unusual in requiring three redox centers to transfer electrons from NADH to the terminal dioxygenase. Although naphthalene dioxygenase is enantiomer-specific in producing the (+)-cis-(1R,2S)dihydrodiol, it also possesses dehydrogenase and monooxygenase activities (Gibson et al. 1995), so that monooxygenase and dioxygenase activities are not exclusive. Exclusive monooxygenation has been observed; for example, the transformation of naphthalene by Bacillus cereus yielded naphth-1-ol with retention of the deuterium after incubation with 1[2H]- and 2-[2H]naphthalene, which is consistent with involvement of an arene oxide and an NIH shift (Cerniglia et al. 1984). Alkylated Naphthalenes Methylnaphthalenes are important components of crude oils, and their degradation follows the initial stages used for alkylated benzenes. The degradation of 2,6-dimethylnaphthalene by flavobacteria involves a pathway analogous to that for dimethylbenzenes—successive oxidation of one methyl group to carboxylate, decarboxylative dioxygenation to 1,2-dihydroxy-6-methylnaphthalene, ring fission to 5-methylsalicylate, followed by further degradation by pathways established for naphthalene itself (Barnsley 1988). Degradation of both 1-methyl- and 2-methylnaphthalene was initiated by 7,8-dioxygenation (Mahajan et al. 1994). Whereas 1-methylnaphthalene produced only 3-methyl catechol, 2-methylnaphthalene produced both 4-hydroxymethylcatechol and 4-methylcatechol. The catechols then underwent extradiol fission. The oxidation of 1,5-, 2,6-, 2,7-, and 1,8-dimethyl naphthalene by a recombinant strain of Pseudomonas aeroginosa PAO1 involved successive oxidation of only a single methyl group to the monocarboxylates, except for 1,8-dimethylnaphthalene in which both methyl groups were oxidized to the dicarboxylate (Selifonov et al. 1996). The oxidation of a wide range of dimethyl naphthalenes has been examined in Sphingomonas paucimobilis strain 2322 (Dutta et al. 1998). Degradative pathways involved successive oxidation to the corresponding

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carboxylate of a methyl group in one of the rings. Further degradation involved the following alternative pathways: a. Decarboxylative dioxygenation to a 1,2-dihydroxynaphthalene followed by 1:1a extradiol fission and formation of salicylate b. Hydroxylation to a 1-hydroxynaphthalene-2-carboxylate and 1:2-intradiol ring fission with the formation of o-phthalate Naphthols and Naphthalene Carboxylates Less attention has been devoted to the degradation of naphthols and naphthalene carboxylates: 1. a. The degradation of naphth-1-ol has been described (Bollag et al. 1975), and the specificity of oxygen uptake in whole cells—though not in cell-free extracts—has been studied (Larkin 1988). Two strains, Pseudomonas sp. strain 12043 and Rhodococcus sp. strain NCIB 12038, were less versatile, whereas Pseudomonas sp. strain 12042 was more versatile, showing oxygen uptake with both 1- and 2-naphthols, 1,5- and 2,7dihydroxynaphthalene and naphthalene itself. The subsequent degradation of the 1,2dihydroxynaphthalene formed from the naphthols may plausibly be assumed to proceed by ring fission and subsequent established reactions that produce salicylate and then catechol (strain 12042) or gentisate (strains 12043 and 12038). b. Naphthalene-1-hydroxy-2-carboxylate is an intermediate in the degradation of phenanthrene, and it has been proposed that as a β-ketoacid, it is decarboxylated to naphth-1-ol followed by pathways leading to salicylate or o-phthalate (Samanta et al. 1999). c. Naphthalene 1,8-dicarboxylate and naphthalene 1-carboxylate are intermediates in the degradation of acenaphthylene by Rhizobium sp. strain CU-A1. They are formed from acenaphthenequinone, and then degraded to salicylate and 2,5-dihydroxybenzoate (Poonthrigpun et al. 2006). 2. Although the degradation of naphthalene-2-carboxylate by Burkholderia sp. strain JT 1500 involves the formation of 1-hydroxy naphthalene-2-carboxylate, this is not formed from the expected (1R,2S)-cis-1,2-dihydrodiol-2-naphthoate. Possibly, therefore, the reaction is carried out by a monooxygenase, or a dehydration step is involved. Subsequent reactions produced pyruvate and o-phthalate that was degraded via 4,5-dihydroxyphthalate (Morawski et al. 1997). Degradation of naphthalene carboxylates formed by oxidation of methyl groups has already been noted. Biphenyl The degradation of biphenyl is known in considerable detail on account of interest in the degradation of PCBs, and virtually all bacteria that degrade PCBs have been isolated after enrichment with biphenyl (Pellizari et al. 1996). Although the degradative pathways of naphthalene and biphenyl are broadly similar, dioxygenation of chlorinated biphenyls by Pseudomonas sp. strain LB400 may occur at either the 2,3- or the 4,5-positions (Haddock et al. 1995). A strain of Pseudomonas paucimobilis, which also degrades toluene, and 1,3- and 1,4-dimethylbenzene, carries out dioxygenation and dehydrogenation of biphenyl followed by ring fission to benzoate, which is further degraded by intradiol ring fission (Figure 8.9) (Furukawa et al. 1983). The three-component biphenyl dioxygenase consists of a reductaseBPH, a ferredoxinBPH, and a terminal dioxygenase ISPBPH. The dioxygenase contains two subunits and a Rieske-type [2Fe–2S] center that are encoded by the genes bphA4 (bphG), bphA3 (bphF), and bphA1bph/A2 (bphA/bphE) (Erickson and Mondello 1992). The dioxygenase has been purified from Pseudomonas sp. strain LB400 (Haddock and Gibson 1995). All the components from Comamonas testosteroni strain B-356, which degrades 4-chlorobiphenyl, have been isolated. Purification of the reductaseBPH and ferredoxinBPH proteins was made possible by using His-tagged components produced from

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OH

H

HO

HO

H

OH O

CO2H

OH CO2H

OH

FIGURE 8.9 Degradation of biphenyl. (From Neilson, A.H. and Allard, A.-S. The Handbook of Environmental Chemistry, Springer, 1998. With permission.)

recombinant strains of Escherichia coli (Hurtubise et al. 1995). The reductaseBPH contains 1 Mol of FAD, the ferredoxinBPH a Rieske [2Fe–2S] center, and together with the terminal oxygenase ISPBPH, both biphenyl-2,3-dihydrodiol and biphenyl-3,4-dihydrodiol were produced from biphenyl. It has been shown that the product of dioxygenation of biphenyl by Sphingomonas yanoikuyae strain B1 (Beijerinckia sp. B1) is biphenyl cis-(2R,3S)-2,3-dihydrodiol (Ziffer et al. 1977). This is then converted into 2,3-dihydroxybiphenyl by a dehydrogenase encoded by the bphB gene followed by ring fission by 2,3-dihydroxybiphenyl 1,2-dioxygenase encoded by the bphC gene. The ring fission dioxygenase has been examined in a number of organisms, and although polyclonal antibodies from the enzymes from P. paucimobilis strain Q1 and from P. pseudoalcaligenes strain KF707 do not cross-react (Taira et al. 1992), there is a high degree of homology between the bphC gene in P. putida strain OU 83 and those from P. cepacia strain LB400 and P. pseudoalcaligenes strain KF707 (Khan et al. 1996). Greater complexity has been observed in Gram-positive strains: in Rhodococcus globerulus strain P6, there are three genes encoding 2,3-dihydroxybiphenyl-1,2dioxygenase (Asturias and Timmis 1993), and in Rhodococcus erythropolis strain TA421, four (Maeda et al. 1995). Although toluene-2,3-dioxygenase, naphthalene-1,2-dioxygenase, and bipheny-2,3-dioxygenase are broadly similar, an interesting difference has emerged in the products formed from benzocycloheptene by bacterial strains that express these activities (Resnick and Gibson 1996). Fungal hydroxylation and subsequent ring fission has only seldom been reported. The imperfect fungus Paecilomyces liliacinus is able to produce successively mono-, di-, and trihydroxylated metabolites from biphenyl and carry out intradiol ring fission of rings carrying adjacent hydroxyl groups (Gesell et al. 2001). PAHs with Three or More Rings The aerobic degradation of a wide range of PAHs with three or more rings has been described, and the pathways are analogous to those outlined for naphthalene and biphenyl—though sometimes they have been covered in less detail. These reactions have been reviewed (Kanaly and Harayama 2000), and only salient features with some exceptions are briefly noted. The flexibility of some organisms—including several mycobacteria—to degrade or metabolize a range of PAHs and the apparently unique metabolic versatility of a strain of Enterobacter sp. is particularly striking. Some illustrative examples are given in Table 8.1.

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TABLE 8.1 Bacterial Degradation of PAHs Substrate Nap, Phe, Flu, 3-Me-Chol Nap, Phe, Anth, Flu Nap, Flu, Pyr Phe, Pyr, Flu Nap, Phe, Anth, Fl Nap, Phe, Anth, Pyr, BaPyr, BaAnth Flu, Pyr, BaAnth, DBaAnth Bi, Phe, Anth, Flu, BbFlu, Chr, Pyr Flu, Pyr, BaAnth, Chr, BaPyr Nap, Anth, Phe, Flu, Pyr, BaAnth, BaPyr Note:

Organism Mycobacterium sp. Alcaligenes denitrificans Enterobacter sp. Mycobacterium sp. BB1 Pseudomonas cepacia F297 Pseudomonas putida GZ44 Mycobacterium sp. RJGII-135 Burkholderia cepacia Pseudomonas paucimobilis Pseudomonas paucimobilis Mycobacterium vanbaalenii PYR-1

Reference Heitkamp et al. (1988a,b) Weissenfels et al. (1991) Sarma et al. (2004) Boldrin et al. (1993) Grifoll et al. (1994, 1995) Goyal and Zylstra (1996) Schneider et al. (1996) Juhasz et al. (1997) Mueller et al. (1990) Ye et al. (1996) Kim et al. (2006)

Bi, biphenyl; Nap, naphthalene; Phe, phenanthrene; Anth, anthracene; Fl, fluoranthene; Pyr, pyrene; Chr, chrysene; BaAnth, benz[a]anthracene; BaPyr, benzo[a]pyrene; BbFlu, benzo[b]fluoranthene; DbaAnth, dibenz[a,h]anthracene; 3-Me-Chol, 3-methylcholanthrene.

The stereospecific dioxygenation of several PAHs by bacteria has been described, and includes the following illustrative examples: a. Anthracene to cis-(1R,2S)-anthracene-1,2-dihydrodiol (Akhtar et al. 1975) b. Phenanthrene to cis-(1R,2S)-phenanthrene-1,2-dihydrodiol and cis-(3S,4R)-phenanthrene3,4-dihydrodiol (Koreeda et al. 1978) c. Benz[a]anthracene to benz[a]anthracene-cis-(1R,2S)-1,2-dihydrodiol, cis-(8R,9S)-8,9dihydrodiol, and cis-(10S,11R)-10,11-dihydrodiol (Jerina et al. 1984) Further degradation of the dihydrodiols involves dehydrogenation to the catechols followed by ring fission and oxidative degradation of the fission product. However, for higher PAHs, degradation may be more complex. Other issues that intrude include dioxygenation at different positions of the rings, and the alternative of monooxygenation. A great deal of attention has been given to the degradation of higher PAHs, and only selected examples are given as illustration. Anthracene and Phenanthrene The degradation of these was described many years ago (Evans et al. 1965), and details have since been added (Figure 8.10a) (van Herwijnen et al. 2003). Phenanthrene is more readily degraded than anthracene, and both can be degraded to o-phthalate. After fission of the peripheral ring in phenanthrene, several pathways have been demonstrated: a. The naphthalene pathway via salicylate (Evans et al. 1965) b. The o-phthalate pathway followed by fission of 3,4-dihydroxybenzoate (Kiyohara and Nagao 1978; Kiyohara et al. 1976; Barnsley 1983) (Figure 8.10b) c. Degradation to diphenyl-6,6′-dicarboxylate (Figure 8.10c) (Moody et al. 2001) d. A pathway that involves naphth-1-ol, which has already been noted (Samanta et al. 1999) The enzymes involved in the degradation of phenanthrene by Nocardioides sp. strain KP7 have been characterized. In this strain, the genes are chromosomal, and phdA and phdB encoding the α- and β-subunits of the dioxygenase, phdC and phdD that encode the ferredoxin, and ferredoxin reductase have been cloned and sequenced (Saito et al. 2000). The order of the genes

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Environmental Degradation and Transformation of Organic Chemicals OH OH

OH

CO2H

CO 2 H

CO2H

(a) OH HO

OH CO2H

CO2H

(b) CO2H (c) CO2H CO2H

OH OH

FIGURE 8.10

Degradation of (a) anthracene, (b,c) alternative pathways for phenanthrene.

O

O

OH H OH

CO2H CO2H OH OH CHO O

O

FIGURE 8.11 Pathways for degradation of fluorene. (From Neilson, A.H. and Allard, A.-S. The Handbook of Environmental Chemistry, Springer, 1998. With permission.) OH

OH OH CHO

O

O

O

FIGURE 8.12 Degradation of fluorene. (From Neilson, A.H. and Allard, A.-S. The Handbook of Environmental Chemistry, Springer, 1998. With permission.)

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was different from that in pseudomonads, and there was only moderate similarity to the sequence for other dioxygenase subunits. 1-Hydroxy-2-naphthoate dioxygenase (Iwabuchi and Harayama 1998a), trans-2′-carboxybenzalpyruvate hydratase-aldolase (Iwabuchi and Harayama 1998b), and 2-carboxybenzaldehyde dehydrogenase, which catalyzes the formation of o-phthalate (Iwabuchi and Harayama 1997), have been purified and characterized. The purified dioxygenase was used to determine further details of the degradation pathways (Adachi et al. 1999). The initial product from ring fission is in equilibrium with the lactone formed by reaction between the carboxyl group and the activated double bond of the benzalpyruvate in a reaction formally comparable to the formation of muconolactones during degradation of monocyclic aromatic compounds. An important exception to the generalization that bacterial dioxygenation is used for aromatic hydrocarbons is provided by Streptomyces flavovirens in which cytochrome P450 produces (–)-trans-[9S,10S]-9,10-dihydrodihydroxyphenanthrene with minor amounts of 9-hydroxyphenanthrene (Sutherland et al. 1990). Fluorene Several pathways have been reported including oxidation to fluoren-9-one and dioxygenation at the ring junction followed by degradation of the oxygenated ring by extradiol fission to yield phthalate (Grifoll et al. 1994; Trenz et al. 1994). In one variant, an 8-hydroxy benzo[c]coumarin is produced (Figure 8.11). Alternatively, dioxygenation may take place to produce 3,4-dihydroxyfluorene followed by extradiol ring fission and further degradation to 3,4-dihydrocoumarin (Grifoll et al. 1992) (Figure 8.12). A strain of Pseudomonas cepacia F297 that was able to grow with a range of PAHs including fluorene, anthracene, and phenanthrene degraded fluorene via 3,4-dihydroxyfluorene and extradiol fission to indan-1-one as terminal metabolite (Grifoll et al. 1994). Fluoranthene Degradation by Mycobacterium sp. strain PYR-1 involves dioxygenation at both the 1,2-, and 7, 8-positions, producing after ring fission fluorene-9-one (Figure 8.13a), or acenaphthene-7-one (Figure 8.13b) (Kelley et al. 1993). In a strain of Alcaligenes denitrificans, acenaphthene-7-one is further

OH

OH (a)

O HO

O

CO2H

O HO (b)

(c)

O O

O

OH

CH2CO2H

FIGURE 8.13 Degradation of fluoranthene. (From Neilson, A.H. and Allard, A.-S. The Handbook of Environmental Chemistry, Springer, 1998. With permission.)

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oxidized by a Baeyer—Villiger-type oxidation to 3-hydroxymethyl-3,4-dihydro-benzo[d,e]coumarin before further oxidation (Weissenfels et al. 1991) (Figure 8.13c). The unusual transformation products fluoranthene-2,3- and 1,5-quinone have been identified in several strains of bacteria (Kazunga et al. 2001). Both were terminal metabolites and since the 2,3-quinone inhibited the degradation of other PAHs, its formation could have an adverse effect on attempts at bioremediation of PAH-contaminated sites. Benz[a]anthracene Transformation by Sphingomonas yanoikuyae strain B1 (Beijerinckia sp. strain B 1) involves dioxygenation at the 1,2-, 8,9-, or 10,11-positions with production of the cis-[1R,2S], cis-[8R,9S], or cis[10S,11R] dihydrodiols of which the first is the dominant (Jerina et al. 1984). These are then further degraded to 1-hydroxy-2-carboxyanthracene, or the corresponding phenanthrenes (Mahaffey et al. 1988) (Figures 8.14a through 8.14c). In contrast, Mycobacterium sp. strain RJGII-135 forms the 5,6- and 10,11-dihydrodiols (Schneider et al. 1996), so that four different pathways for degradation are possible. Exceptionally, monooxygenation may occur concomitant with dioxygenation; for example, the transformation of 7,12-dimethylbenz[a]anthracene by Mycobacterium vanbaalenii PYR-1 takes place by monooxygenation to epoxides followed by hydration to trans-dihydrodiols, and also by dioxygenation to cis-dihydrodiols (Moody et al. 2003). Pyrene Degradation by a Mycobacterium sp. involves both monooxygenation and dioxygenation at the 4,5-position with production of the trans- and cis-dihydrodiols, respectively, which are degraded to phenanthrene-4-carboxylate and ultimately to o-phthalate. 4-Hydroxyperinaphthenone was also produced, presumably by initial dioxygenation at the 1,2-positions (Heitkamp et al. 1988a) (Figure 8.15). This is consistent with the structure of the intermediate from the degradation of pyrene by Rhodococcus sp. strain UW1 (Walter et al. 1991). Alternatively, degradation by Mycobacterium sp. OH HO

OH CO2H

(a)

CO2H OH

(b) HO OH OH (c) OH HO CO2H

FIGURE 8.14 Pathways for partial degradation of benz[a]anthracene. (From Neilson, A.H. and Allard, A.-S. The Handbook of Environmental Chemistry, Springer, 1998. With permission.)

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HO2C CO2H OH

CO2H

OH OH O

OH

OH

FIGURE 8.15

Degradation of pyrene.

OH

OH

OH

OH CO2H + CO2H CO2H

CO2H CO2H HO OH

CO2H CO2H

FIGURE 8.16

Degradation of pyrene.

strain AP1 produced both cis- and trans-4,5-dihydrodiols (Vila et al. 2001). The former was degraded further to phenanthrene-4,5-dicarboxylic acid, phenthanrene-4-carboxylate and phthalate, and 6,6′-dihydroxy-biphenyl-2,2′-dicarboxylate (Figure 8.16). Bacterial formation of pyrene-4,5quinone has been reported (Kazunga and Aitken 2000). Although the degradation of pyrene by Mycobacterium sp. strain A1-PYR was restricted with the formation of several oxygenated metabolites, complete degradation without their formation was accomplished when phenanthrene or fluoranthene was also present (Zhong et al. 2006). This is particularly relevant to the degradation of mixtures of PAHs, for example, in bioremediation (Chapter 14, Part 2).

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OH

CO2H CO2H

OH

FIGURE 8.17 Products produced by partial degradation of benzo[a]pyrene. (From Neilson, A.H. and Allard, A.-S. The Handbook of Environmental Chemistry, Springer, 1998. With permission.)

Benzo[a]pyrene Mycobacterium sp. strain RJGH-135, which was able to metabolize pyrene, benz[a]anthracene, and benzo[a]pyrene, degraded the last by dioxygenation at the 4,5-, 7,8-, and 9,10-positions (Schneider et al. 1996). Chrysene-4,5-dicarboxylate was presumptively formed by intradiol fission of the 4, 5-benzo[a]pyrene-4,5-diol, while the 7,8- and 9,10-dihydrodiols underwent extradiol fission to 7- and 8-dihydropyrene carboxylates (Figure 8.17). Mycobacterium vanbaalenii strain PYR1 carried out both monooxygenation with the formation of benzo[a]pyrene trans-11,12-dihydrodiol and dioxygenation to cis-dihydrodiols at the 4,5- and 11,12-positions. The former was then degraded further to chrysene-4,5-dicarboxylic acid (Moody et al. 2004). A soil strain of Rhodanobacter sp. in a consortium could degrade [7-14C]benzo[a]pyrene apparently using metabolites produced by other organisms, since the pure cultures were unable to degrade the substrate (Kanaly et al. 2002). This may be a widespread phenomenon for complex mixtures such as those that were used in that study. Anaerobic Reactions Carried Out by Bacteria The anaerobic degradation of a few PAHs has been examined, generally using mixed cultures from widely different habitats, and investigations have been limited hitherto to naphthalene and phenanthrene. Although no pathway was established, the degradation has been assessed from the diminishing substrate concentration of anthracene, phenanthrene, and pyrene by denitrifying pseudomonads isolated from diverse environments under both aerobic and anaerobic denitrifying conditions (McNally et al. 1998). Under sulfate-reducing conditions, [1-14C]naphthalene and [9-14C]phenanthrene were oxidized to 14CO2 (Coates et al. 1996, 1997), and a pure culture NaphS2b that belongs to the δ-subclass of the Proteobacteria has been isolated, which can degrade naphthalene, naphthalene-2-carboxylate, benzoate, and acetate (Galushko et al. 1999). Under anaerobic nitrate-reducing conditions, both [ring 14C]toluene and [1-14C]naphthalene were oxidized to 14CO2 (Bregnard et al. 1996), and pure cultures of denitrifying organisms could mineralize [U14C]naphthalene under anaerobic conditions, although a substantial fraction of the label was recovered in cell biomass (Rockne et al. 2000). Although carboxylated intermediates of naphthalene and phenanthrene have been identified in a sulfidogenic mixed culture (Zhang and Young 1997), these are not the primary metabolites. Details of the pathway emerged using a sulfate-reducing enrichment culture that was able to grow at the expense of naphthalene and 2-methylnaphthalene. It was shown (Safinowski and Meckenstock 2006) that transformation of naphthalene is initiated by the formation of 2-methylnaphthalene, which is able to support the growth of this enrichment culture. Its transformation is initiated by reaction with fumarate analogous to that used for the activation of toluene. The pathway involves successive formation of naphthyl-2-methylsuccinate, naphthyl-2-methylenesuccinate, and naphth2-oate, followed by subsequent steps to 2-carboxycyclohexylacetate (Figure 8.18) (Annweiler et al. 2002) by the reduction of both rings (Meckenstock et al. 2000) and hydroxylation; these steps are analogous to those for benzoate. In cometabolism with 2-methylnaphthalene, analogous

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CH2

CO 2 H

409 CO2H

CO2H

CO 2 H

CO2H OH

FIGURE 8.18

CO2H CO2H

CO2H CO2H

Anaerobic degradation of 2-methylnaphthalene.

reactions were shown to operate with benzothiophene and benzofuran (Safinowski et al. 2006). It was suggested that these polar metabolites could be used to assess degradation of these arenes in anaerobic groundwater, analogous to the use of benzylsuccinates from monocyclic arenes (BTEX), which has already been demonstrated (Beller et al. 1995, 2002). An analogous pathway that produced decahydronaphthalene-2-carboxylate has been demonstrated for another sulfate-reducing consortium in which reduction of the A-ring and carboxylation of the B-ring of 2-methylnaphthalene was also observed (Sullivan et al. 2001). Fungal Transformations In contrast to the degradations carried out by bacteria, eukaryotic organisms such as yeasts and fungi often accomplish only biotransformation. The aryl hydrocarbon monooxygenase in Cunninghamella baineri is a cytochrome P450 (Ferris et al. 1976) in which the [FeO]3+complex with the substrate may carry out either direct hydroxylation of the substrate by elimination of a proton, or produce an epoxide or a rearranged (NIH shift) monohydroxy (phenolic) compound (Guengerich 1990). Isotope experiments clearly show that the phenols are generally not produced by either a direct substitution reaction or elimination from the trans-dihydrodiols. Epoxide hydrolases produce phenols from the trans-dihydrodiols, and can then be conjugated to sulfate or glucuronide esters (Cerniglia et al. 1982b), or O-methylated. Monooxygenation of methyl groups to carboxyl groups has also been reported. C. elegans has been used to examine the transformation of a number of polycyclic aromatic hydrocarbons. Reactions are generally confined to oxidation of the rings with the formation of phenols, catechols, and quinones, and ring fission does not generally take place. Different rings may be oxygenated, for example, in 7-methylbenz[a]anthracene (Cerniglia et al. 1982c) (Figure 8.19) or oxidation may take place in several rings, for example, in fluoranthene (Pothuluri et al. 1990) (Figure 8.20). Since fungal metabolism of PAHs has been suggested as a model for mammalian metabolism (Smith and Rosazza 1983), particular attention has been directed to the stereochemistry and absolute configuration of the trans-dihydrodiols produced by monooxygenation. There is one very significant difference between the trans-dihydrodiols produced by fungi and those from mammalian systems—the absolute configuration of the products. Although trans-1,2-dihydroxy-1, 2-dihydroanthracene and trans-1,2-dihydroxy-1,2-dihydrophenanthrene are formed from anthracene by C. elegans, these dihydrodiols have the S,S configuration in contrast to the R,R configuration of the metabolites from rat liver microsomes (Cerniglia and Yang 1984). It has become clear, however, that the situation among a wider range of fungi is much less straightforward. For example, the trans-9,10-dihydrodiol produced by Phanerochaete chrysosporium was predominantly the 9S,10S enantiomer, whereas those produced by Cunninghamella elegans and Syncephalastrum racemosum were dominated by the 9R,10R enantiomers (Sutherland et al. 1993). Comparable differences were also observed for the trans-1,2-dihydrodiols and trans-3,4-dihydrodiols, so that generalizations on the enantiomeric selectivity of these reactions should be viewed with caution.

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CH2OH

CH3 HO H HO H

FIGURE 8.19

CH2OH

Biotransformation of 7-methylbenz[a]anthracene by Cunninghamella elegans.

OH HO

H

H

HO

OH HO

FIGURE 8.20

H

H

Alternative pathways for the biotransformation of fluoranthene by Cunninghamella elegans.

There are two issues that recur in fungal transformations, and these have been extensively documented: 1. The phenol, which is formed by rearrangement from the initially produced transdihydrodiol, may be conjugated to form sulfate esters or glucuronides (Cerniglia et al. 1982b; Golbeck et al. 1983; Cerniglia et al. 1986; Lange et al. 1994). The less-common glucosides have also been identified: 1-phenanthreneglucopyranoside is produced from phenanthrene by Cunninghamella elegans (Cerniglia et al. 1989) and 3-(8-hydroxyfluoranthene)lucopyranoside from fluoranthene by the same organism (Pothuluri et al. 1990). The xylosylation of 4-methylguaiacol and vanillin by the basidiomycete Coriolus versicolor (Kondo et al. 1993) represents an even less-common alternative. 2. In reactions involving monooxygenase systems with the formation of intermediate arene oxides, rearrangement of substituents may take place (Figure 8.21a). This is an example of the NIH shift that plays an important role in the metabolism of xenobiotics by mammalian systems (Daly et al. 1972). It has also been observed in fungal (Figure 8.21b) (Faulkner and Woodcock 1965; Smith et al. 1981; Cerniglia et al. 1983) and bacterial systems

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R

R H

(a) x

CH2−CO2H O

O Cl

411

R X

O

H X

X

O

OH

CH2−CO2H

O

Cl

CH2−CO2H Cl

(b) HO Cl

Cl Cl

CH2−CH3

OH

CH2−CH3

(c) H D

FIGURE 8.21

D H

Examples of the NIH shift.

(Figure 8.21c) (Dalton et al. 1981; Cerniglia et al. 1984a,b; Adriaens 1994) including the marine cyanobacterium Oscillatoria sp. (Narro et al. 1992a,b). The transformation of naphthalene by a large number of fungi has been examined, including representatives of the Mucorales such as Cunninghamella, Syncephalastrum, and Mucor, which were the most active. The principal product was naphth-1-ol, with lesser amounts of 2-naphthol, 4-hydroxytetral-1-one, trans-naphthalene-1,2-dihydrodiol, and the 1,2- and 1,4-quinones (Cerniglia et al. 1978). The results of detailed study (Cerniglia et al. 1983) of the transformation of naphthalene by Cunninghamella elegans illustrate most of the basic principles involved in the fungal transformation of all PAHs: a. The initial product is naphthalene-1,2-epoxide, which is converted into the trans-dihydrodiol with the (+)-(1S,2S) configuration. b. The trans-dihydrodiol is formed by introduction of a single atom of 18O2 at C1 while that at C2 comes from H218O. c. The naphth-1-ol formed from [1-2H]-naphthalene retains 78% of the deuterium so that an NIH shift is involved. Similar experiments with 1-methyl- and 2-methylnaphthalene involved oxidation of the methyl group to hydroxymethyl and carboxyl groups (Cerniglia et al. 1984a,b). The transformation of biphenyl is analogous to that of naphthalene, with the production of hydroxylated biphenyls as the major metabolites: 4-hydroxybiphenyl from C. echinulata, 2-hydroxybiphenyl from Helicostylum piriforme, 4,4′-dihydroxybiphenyl from Aspergillus parasiticus (Golbeck et al. 1983; Smith et al. 1980), and 2-, 3-, and 4-hydroxybiphenyl, and 4,4′-dihydroxybiphenyl from C. elegans (Dodge et al. 1979). During the formation of 4-hydroxybiphenyl from [U-2H]-biphenyl by C. echinulata, 20% of the deuterium was retained, which is consistent with formation of an arene oxide followed by an NIH shift (Smith et al. 1981).

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H

OH

H HO HO HH

OH O

O

O

O

FIGURE 8.22 Products produced by fungal transformation of benzo[a]pyrene. (From Neilson, A.H. and Allard, A.-S. The Handbook of Environmental Chemistry, Springer, 1998. With permission.)

Higher PAHs may produce a number of products. For example, 7,12-dimethylbenz[a] anthracene gives rise to a range of metabolites including both hydroxymethyl compounds and several trans-dihydrodiols. The 7-hydroxymethyl-12-methylbenz[a]anthracene-trans-5,6-dihydrodiols produced by C. elegans and S. racemosum were mixtures of the R,R and S,S enantiomers with the former predominating (McMillan et al. 1987). This is, however, contrary to the products from mammalian systems. From fluoranthene, C. elegans produces both the trans-2,3-dihydrodiol as well as the 8-hydroxy-3-hydroxyglucoside, and the 8- and 9-hydroxy-3,4-dihydrodiols (Pothuluri et al. 1990). The metabolism of pyrene and benzo[a]pyrene by C. elegans is increasingly complex. Pyrene is transformed by hydroxylation at the 1-, 1- and 6-, and 1- and 8- positions, and the bisphenols were glucosylated at the 6- and 8-positions; the 1,6- and 1,8-pyrenequinones were also formed (Cerniglia et al. 1986). The same organism transformed benzo[a]pyrene to the trans-7,8- and trans9,10-diols, the 3- and 9-hydroxy compounds, and the 1,6- and 3,6-quinones (Cerniglia and Gibson 1979) (Figure 8.22). In addition, C. elegans formed sulfate conjugates of polar metabolites, and one of these was identified as the 7b,8a,9a,10b-tetrahydrotetrol formed sequentially from the 7b,8bepoxide, the 7b,8a-dihydrodiol, and the 7b,8a-dihydrodiol-9a,10a-epoxide (Cerniglia and Gibson 1980a). Tetrahydrotetrols were also formed by metabolism of both diastereomers of (+/–)-trans9,10-dihydrodiol including the 7b,8a,9b,10a- and 7b,8a,9a,10b-compounds; a dihydrodiol epoxide was also isolated that could be hydrolyzed to the former (Cerniglia and Gibson 1980b). The transformation of 3-methylcholanthrene by C. elegans was exceptional in that 1-hydroxy-trans-9,10dihydrodiol was formed in only trace amounts, and benzylic oxidations at C1 and C2 dominated (Cerniglia et al. 1982d). The metabolism of pyrene by the basidiomycete fungus Crinipellis stipitaria strain JK 364 illustrates a hitherto largely neglected possibility: biosynthesis from the products of metabolism. This strain can transform pyrene to 1-hydroxypyrene, and trans-4,5-dihydro-4,5-dihydroxypyrene. Simultaneously, the secondary metabolite 3-methyl-6,8-dihydroxyisocoumarin is produced, which is synthesized from acetate produced by ring fission (Lange et al. 1995). The biosynthesis of the isocoumarin was also induced by chrysene, though not by anthracene, phenanthrene, or fluoranthene. Most of the foregoing studies were conducted with C. elegans or S. racemosum, but results have been presented showing that a wide range of nonbasidiomycete soil fungi are capable of oxidizing both pyrene and benzo[a]pyrene (Launen et al. 1995). Those with high activity included Penicillium janthinellum that produced 1-hydroxypyrene as a major metabolite, together with the 1,6- and 1,8-quinones.

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Yeasts and Algae A cytochrome P450 has been purified from Saccharomyces cerevisiae that has benzo[a]pyrene hydroxylase activity (King et al. 1984), and metabolizes benzo[a]pyrene to 3- and 9hydroxybenzo[a]pyrene and benzo[a]pyrene-7,8-dihydrodiol (Wiseman and Woods 1979). The transformation of PAHs by Candida lipolytica produced predominantly monohydroxylated products: naphth-1-ol from naphthalene, 4-hydroxybiphenyl from biphenyl and 3- and 9hydroxybenzo[a]pyrene from benzo[a]pyrene (Cerniglia and Crow 1981). The transformation of phenanthrene was demonstrated in a number of yeasts isolated from littoral sediments and of these, Trichosporum penicillatum was the most active. In contrast, biotransformation of benz[a]anthracene by Candida krusei and Rhodotorula minuta was much slower (MacGillivray and Shiaris 1993). Naphthalene, biphenyl, and phenanthrene may be transformed in low yield by oxygenic phototrophs (algae and cyanobacteria) with the formation of several metabolites. A range of algae produced naphth-1-ol from naphthalene (Cerniglia et al. 1980a,b; 1982a), while biphenyl was transformed by a strain of the cyanobacterium Oscillatoria sp. to 4-hydroxybiphenyl (Cerniglia et al. 1984a). A study with the same strain showed that the naphth-1-ol formed from [1-2H]- and [2-2H]naphthalene retained 68% and 74% of the deuterium. This shows that the reaction proceeded by initial formation of naphthalene-1,2-epoxide followed by an NIH shift (Narro et al. 1992a). The transformation of phenanthrene by the unicellular cyanobacterium Agmenellum quadruplicatum involved monooxygenation to produce the trans 9,10-dihydrodiol and lesser amounts of 3-methoxyphenanthrene. The dihydrodiol which has an enantiomeric excess of the (–)-[9S,10S] enantiomer was produced by the introduction of a single atom of 18O2 (Narro et al. 1992b), and is formally analogous to the transformation of naphthalene to naphth-1-ol by a strain of Oscillatoria sp., which has already been noted. The products from the metabolism of benzo[a]pyrene by the green alga Selenastrum capricornutum strain UTEX 1648 are unexpected, since the cis-dihydrodiols at the 11,12-, 7,8-, and 4,5-positions were apparently produced by dioxygenation (Warshawsky et al. 1995b). The light conditions determine both the preferred site of dioxygenation and the relative contribution of the phytotoxic benzo[a]pyrene-3,6-quinone (Warshawsky et al. 1995a). An interesting and exceptional example of ring fission by algae is afforded by the degradation of phenol by Ochromonas danica via catechol followed by extradiol fission (Semple and Cain 1996). White-Rot Fungi There has been considerable interest in white-rot fungi since they are capable of degrading not only lignin but also a wide range of other substrates including PAHs. They are discussed separately since the mechanisms whereby they accomplish degradation or transformation of PAHs differ significantly from the fungi already discussed. Although considerable attention has been directed to Phanerochaete chrysosporium, the degradation and transformation of PAHs is not limited to this taxon (Ferris et al. 1976). Investigations with Ph. laevis and Pleurotus ostreatus (Bezalel et al. 1996a) have brought to light significantly different metabolic pathways. It should be pointed out that ten different strains of Ph. chrysosporium have been shown to harbor bacteria, although only a few of these have been identified (Seigle-Murandi et al. 1996). The metabolic consequence of this association is apparently unresolved. The mechanism and enzymology of transformations by white-rot fungi are complicated for several issues: a. The synthesis of the lignin and Mn peroxidases is regulated by several factors including the nitrogen status of the cells, the oxygen concentration, and the concentration of Mn2+ in the growth medium. b. The dependence of enzyme activity on the growth phase of the cells. The results of experiments with whole cells may, therefore, differ from those obtained with pure enzymes or cell extracts. c. The range of oxygenation mechanisms that has emerged.

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A brief summary of these factors includes the following issues: 1. Peroxidase systems. There are at least three of these extracellular enzyme systems: (a) lignin peroxidase (LiP), (b) manganese-dependent peroxidase (MnP), and (c) manganese peroxide-dependent lipid peroxidase. The first functions by formation of a cation radical that undergoes further reaction, the second by formation of Mn3+ that brings about oxidation of the substrate, and the third probably by proton abstraction followed by the introduction of oxygen. The enzyme system responsible for lignin degradation is expressed during idiophasic growth as a result of nitrogen limitation. The regulation of the synthesis of LiP and MnP is complex, and depends on the Mn2+ concentration in the medium (Bonnarme and Jeffries 1990; Brown et al. 1990), although it is not apparently subject to carbon substrate regulation (van der Woude et al. 1993). However, the synthesis of manganese peroxidase is regulated by concentrations of H2O2 and O2 (Li et al. 1995). 2. Monooxygenases. Under nonlignolytic conditions, arene monooxygenase and epoxide hydrolase systems may function to produce trans-dihydrodiols. Hydrogen abstraction mediated by the lipid peroxidase system may operate, for example, in the formation of fluorene-9-one from fluorene by Ph. chrysosporium (Bogan et al. 1996). 3. Hydroxylases. Some substrates are activated to cation radicals that react further with H2O to ultimately produce quinones. This is particularly prevalent in 4- and 5-ring compounds, though important alternatives exist. 4. Dioxygenase. Although it is not involved in the transformation of PAHs, it may be noted for the sake of completeness that a catechol intradiol dioxygenase is involved in the fission of 1,2,4-trihydroxybenzene that is formed from a number of aromatic substrates (Rieble et al. 1994). Although a number of white-rot fungi have been examined and shown to degrade PAHs (Field et al. 1992), greatest attention has probably been directed to Phanerochaete chrysosporium and Pleurotus ostreatus, and to the PAHs anthracene, phenanthrene, pyrene, and benzo[a]pyrene that will be used to illustrate the cardinal principles. A substantial fraction of PAHs may also be sorbed to the biomass—40% for phenanthrene and 22% for benzo[a]pyrene (Barclay et al. 1995). The degree of mineralization of PAHs by white-rot fungi may sometimes be quite low, for example, for Pleurotus ostreatus, yields were 3.0, 0.44, 0.19, and 0.19% for phenanthrene, pyrene, fluorene, and benzo[a]pyrene, respectively (Bezalel et al. 1996a). Anthracene Degradation by Ph. chrysosporium took place by initial formation of anthra-9,10-quinone followed by ring fission to phthalate (Figure 8.23) (Hammel et al. 1991). This pathway is completely different from that used by both bacteria and fungi, and the fission of the quinone might plausibly involve a Baeyer–Villiger-type insertion of oxygen. In contrast, P. ostreatus formed both anthra-9,10-quinone and anthracene trans-1,2-dihydrodiols as terminal metabolites, with an S,S- to R,R- ratio of 58:42 in the latter (Bezalel et al. 1996c).

O CO2H

O

CO2H

FIGURE 8.23 Degradation of anthracene by Phanerochaete chrysosporium. (From Neilson, A.H. and Allard, A.-S. The Handbook of Environmental Chemistry, Springer, 1998. With permission.)

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CO2H CO2H

O O

OH H OH H

+ H OH

3-ol+4-ol

H

OH

9-ol

FIGURE 8.24 Alternative pathways for transformation of phenanthrene by Phanerochaete chrysosporium. (From Neilson, A.H. and Allard, A.-S. The Handbook of Environmental Chemistry, Springer, 1998. With permission.)

Phenanthrene There are different pathways that have been delineated and these are considered separately as follows: 1. The transformation of phenanthrene has been extensively studied and illustrates the operation of alternative pathways that depend on the status of the cells. During cell growth of Ph. chrysosporium, the trans-3,4- and trans-9,10-diols were formed and rearranged to produce 3-, 4-, and 9-phenanthrenols (Figure 8.24). Since lignin peroxidase activity was not observed, it was suggested that the trans-diols were formed by monooxygenase and epoxide hydrolase activities (Sutherland et al. 1991). 2. In conditions under which lignin degradation is induced, phenanthrene is oxidized by cell suspensions of Ph. chrysosporium to phenanthrene-9,10-quinone, which is then fissioned to biphenyl-2,2′-dicarboxylate (Hammel et al. 1992) (Figure 8.24). Cytochrome P450 activation was not apparently involved, and extracellular ligninases were inactive in vitro. The same products, with a preponderance of the intermediate trans-[R,R]-9,10-dihydrodiol, were produced by P. ostreacus, and it was suggested that a cytochrome P450 system was involved (Bezalel et al. 1996b, 1997). 3. Biphenyl-2,2′-dicarboxylate was formed by Ph. chrysosporium in a reaction mixture with manganese peroxidase, O2, and unsaturated lipid, and it was suggested that a MnPmediated lipid peroxidation was involved (Moen and Hammel 1994). A synthesis of these results has been presented, and evidence presented for the role of alkoxy radicals generated during lipid oxidation (Bogan and Lamar 1995; Tatarko and Bumpus 1993). Pyrene and Benzo[a]pyrene Purified ligninase H8 produced by P. chrysosporium in stationary cultures oxidized pyrene to pyrene-1,6- and pyrene-1,8-quinones in high yield, and experiments with H218O showed that both quinone oxygen atoms originated in water (Figure 8.25). It was suggested that initial one-electron abstraction produced cation radicals at the 1 and 6 or 8-positions (Hammel et al. 1986), whereas in

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O

+

O

O

O

O

+

O

O

FIGURE 8.25 Quinones produced from pyrene and benzo[a]pyrene. (From Neilson, A.H. and Allard, A.-S. The Handbook of Environmental Chemistry, Springer, 1998. With permission.)

P. ostreacus that produced the trans-4,5-dihydrodiols with a 63:37 ratio of the R,R and S,S enantiomers, a cytochrome P450 monooxygenase was involved (Bezalel et al. 1996b). In analogy with the transformation of pyrene, the 1,6-, 3,6-, and 6,12-quinones were formed from benzo[a]pyrene by Ph. chrysopsporium (Haemmerli et al. 1986). These reactions are analogous to those involved in the electrochemical oxidation of benzo[a]pyrene (Jeftic and Adams 1970), and oxidation by Mn(III) acetate (Cremonesi et al. 1989), while substantial evidence supports the role of cation radicals as intermediates in biological oxidations and the tumorogenic properties of some PAHs (Cavalieri and Rogan 1998). In most of the above illustrations, quinones have been formed and they may be terminal metabolites or only transient intermediates that are produced, for example, during mineralization by the basiomycete Stropharia coronillla (Steffen et al. 2003). Degradation was stimulated by addition of Mn2+, and could be attributed to elevated levels of the lignolytic manganese peroxidase. Experiments with Ph. laevis strain HHB-1625 have revealed some additional features (Bogan and Lamar 1996): a. Under conditions of nitrogen limitation in the presence of Mn2+, manganese peroxidase activity was induced although no lignin peroxidase activity could be demonstrated. b. Transformation of anthracene, phenanthrene, benz[a]anthracene, and benzo[a]pyrene was demonstrated without the accumulation of quinones that were, at best, transient metabolites. c. In comparison with Ph. chrysosporium, Ph. laevis was more effective in mineralizing [5,6-14C]benz[a]anthracene-7-12-quinone. In summary, it may be stated that for a wide range of PAHs, the specific association of dioxygenation to give cis-dihydrodiols by prokaryotes and of monooxygenation to epoxides and thence to trans-dihydrodiols by eukaryotes is generally valid. The absolute stereochemistry of the transdihydrodiols produced by fungi is highly variable, and there are frequently significant differences between the products of fungal monooxygenation and those carried out by rat-liver microsomes. Although quinone are frequently formed as metabolites of the white-rot fungus Ph. chrysosporium, these are only transiently formed by Ph. laevis, so that care should be exercised in making generalizations on the basis the results obtained for a single species or strain. The intermediate epoxides may rearrange to produce phenols that are generally conjugated.

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Brown JA, JK Glenn, MA Gold (1990) Manganese regulates expression of manganese peroxidase by Phanerochaete chrysosporium. J Bacteriol 172: 3125–3130. Cavalieri E, E Rogan (1998) Mechanisms of tumor initiation by polycyclic aromatic hydrocarbons in mammals. Handbook Environ Chem 3J: 82–117. Cerniglia CE, DT Gibson (1979) Oxidation of benzo[a]pyrene by the filamentous fungus Cunninghamella elegans. J Biol Chem 254: 12174–12180. Cerniglia CE, DT Gibson (1980a) Fungal oxidation of benzo[a]pyrene and (+/–)-trans-7,8-dihydroxy-7,8dihydrobenzo[a]pyrene: evidence for the formation of a benzo[a]pyrene 7,8-diol-9,10-epoxide. J Biol Chem 255: 5159–5163. Cerniglia CE, DT Gibson (1980b) Fungal oxidation of (+/–)-9,10-dihydroxy-9-10-dihydrobenzo[a]pyrene: formation of diastereomeric benzo[a]pyrene 9,10-diol-7,8-epoxides. Proc Natl Acad USA 77: 4554–4558. Cerniglia CE, DT Gibson, C van Baalen (1980a) Oxidation of naphthalene by cyanobacteria and microalgae. J Gen Microbiol 116: 495–500. Cerniglia CE, DW Kelley, JP Freeman, DW Miller (1986) Microbial metabolism of pyrene. Chem-Biol Interactions 57: 203–216. Cerniglia CE, DT Gibson, C van Baalen (1982a) Naphthalene metabolism by diatoms isolated from the Kachemak Bay region of Alaska. J Gen Microbiol 128: 987–990. Cerniglia CE, JP Freeman, FE Evans (1984a) Evidence for an arene oxide-NIH shift pathway in the transformation of naphthalene to 1-naphthol by Bacillus cereus. Arch Microbiol 138: 283–286. Cerniglia CE, JP Freeman, RK Mitchum (1982b) Glucuronide and sulfate conjugation in the fungal metabolism of aromatic hydrocarbons. Appl Environ Microbiol 43: 1070–1075. Cerniglia CE, JR Althus, FE Evans, JP Freeman, RK Mitchum, SK Yang (1983) Stereochemistry and evidence for an arene oxide-NIH shift pathway in the fungal metabolism of naphthalene. Chem-Biol Interactions 44: 119–132. Cerniglia CE, KL Lambert, DW Mille, JP Freeman (1984b) Transformation of 1- and 2-methylnaphthalene by Cunninghamella elegans. Appl Environ Microbiol 47: 111–118. Cerniglia CE, PP Fu, SK Yang (1982c). Metabolism of 7-methylbenz[a]anthracene and 7-hydroxymethylbenz [a]anthracene by Cunninghamella elegans. Appl. Environ. Microbiol. 44: 682–689. Cerniglia CE, RH Dodge, DT Gibson (1982d) Fungal oxidation of 3-methylcholanthrene: formation of proximate carcinogenic metabolites of 3-methylcholanthrene. Chem-Biol Interactions 38: 161–173. Cerniglia CE, RL Herbert, PJ Szaniszlo, DT Gibson (1978) Fungal metabolism of naphthalene. Arch Microbiol 117: 135–143. Cerniglia CE, SA Crow (1981) Metabolism of aromatic hydrocarbons by yeasts. Arch Microbiol 129: 9–13. Cerniglia CE, SK Yang (1984) Stereoselective metabolism of anthracene and phenanthrene by the fungus Cunninghamella elegans. Appl. Environ. Microbiol. 47: 119–124. Cerniglia CE, C van Baalen, DT Gibson (1980b) Oxidation of biphenyl by the cyanobacterium, Oscillatoria sp. strain JCM. Arch Microbiol 125: 203–207. Cerniglia CE, WL Campbell, JP Freeman, FE Evans (1989) Identification of a novel metabolite in phenanthrene metabolism by the fungus Cunninghamella elegans. Appl Environ Microbiol 55: 2275–2279. Coates JD, J Woodward, J Allen, P Philip, DR Lovley (1997) Anaerobic degradation of polycyclic aromatic hydrocarbons and alkanes in petroleum-contaminated marine harbor sediments. Appl Environ Microbiol 63: 3589–3593. Coates JD, RT Anderson, DR Lovley (1996) Oxidation of polycyclic aromatic hydrocarbons under sulfatereducing conditions. Appl Environ Microbiol 62: 1099–1101. Cremonesi P, EL Cavalieri, EG Rogan (1989) One-electron oxidation of 6-substituted benzo[a]pyrenes by manganic acetate : a model for metabolic activation. J Org Chem 54: 3561–3570. Dalton H, BT Gording, BW Watyers, R Higgins, JA Taylor (1981) Oxidations of cyclopropane, methylcyclopropane, and arenes with the monooxygenase system from Methylococccus capsulatus. J Chem Soc Chem Commun 482–483. Daly JW, DM Jerina, B Witkop (1972) Arene oxides and the NIH shift: the metabolism, toxicity and carcinogenicity of aromatic compounds. Experientia 28: 1129–1149. Demanèche S, C Meyer, J Micoud, M Louwagie, JC Wilson, Y Jouanneau (2004) Identification and functional analysis of two ring-hydroxylating dioxygenases from a Sphingomonas strain that degrades various polycyclic aromatic hydrocarbons. Appl Environ Microbiol 70: 6714–6725. Dodge RH, CE Cerniglia, DT Gibson (1979) Fungal metabolism of biphenyl. Biochem J 178: 223–230. Dunn NW, HM Dunn, RA Austen (1980) Evidence for the existence of two catabolic plasmids coding for the degradation of naphthalene. J Gen Microbiol 117: 529–533.

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Warshawsky D, T Cody, M Radike, R Reilman, B Schujann, K LaDow, J Schneider (1995b) Biotransformation of benzo[a]pyrene and other polycyclic aromatic hydrocarbons and heterocyclic analogues by several green algae and other algal species under gold and white light. Chem-Biol Interactions 97: 131–148. Weissenfels WD, M Beyer, J Klein, HJ Rehm (1991) Microbial metabolism of fluoranthene: isolation and identification of fission products. Appl Microbiol Biotechnol 34: 528–535. Wiseman A, LFJ Woods (1979) Benzo[a]pyrene metaboltes formed by the action of yeast cytochrome P-450/ P-448. J Chem Tech Biotechnol 29: 320–324. Ye D, MA Siddiqui, AE Maccubbin, S Kumar, HC Sikka (1996) Degradation of polynuclear aromatic hydrocarbons by Sphingomonas paucimobilis. Environ Sci Technol 30: 136–142. Yen K-M, CM Serdar (1988) Genetics of naphthalene catabolism in pseudomonads. CRC Crit Rev Microbiol 15: 247–268. Zhang X, LY Young (1997) Carboxylation as an initial reaction in the anaerobic metabolism of naphthalene and phenanthrene by sulfidogenic consortia. Appl Environ Microbiol 63: 4759–4764. Zhong Y, T Luan, H Zhou, C Lan, N Fung, FY Tam (2006) Metabolite production in degradation of pyrene alone or in mixture with another polycyclic aromatic hydrocarbon by Mycobacterium sp. Environ Toxicol Chem 25: 2853–2859. Zhou N-Y, J Al-Dulayymi, MS Baird, PA Williams (2002) Salicylate 5-hydroxylase from Ralstonia sp. strain U2; a monooxygenase with close relationships to and shared electron transport proteins with naphthalene dioxygenase. J Bacteriol 184: 1547–1555. Ziffer H, K Kabuta, DT Gibson, VM Kobal, DM Jerina (1977) The absolute stereochemistry of several cis dihydrodiols microbially produced from substituted benzenes. Tetrahedron 33: 2491–2496.

PART 3 AROMATIC CARBOXYLATES, CARBOXALDEHYDES, AND RELATED COMPOUNDS INTRODUCTION Catechols, benzoates, and hydroxylated benzoates are intermediates in the aerobic degradation of aromatic hydrocarbons. The first stage in their degradation is the introduction of the elements of dioxygen into the ring, either both atoms by dioxygenation or only one atom by monooxygenation. The second stage involves fission of the ring by another group of dioxygenases. Details of these oxygenases are given in Chapter 3, Part 1. The anaerobic degradation of benzoate is carried out by different reactions involving hydrogenation of the ring.

BENZOATES Aerobic Conditions Mechanisms for the Initial Oxygenation Benzoate is generally degraded to catechol by dioxygenation with loss of the carboxyl group. For benzoates substituted at the ortho position with amino or halogen groups, dioxygenation involves the loss of both the carboxyl and amino groups or both the carboxyl and halogen groups. In contrast to arene hydrocarbons, dehydrogenation is not, therefore, required for the formation of catechols. The formation of 1-hydroxynaphthalene-2-carboxylate from naphthalene-2-carboxylate was apparently not via (1R,2S)-cis-1,2-dihydro-1,2-dihydroxynaphthalene-2-carboxylate (Morawski et al. 1997). This is supported by the activity of naphthalene dioxygenase in which both activities are catalyzed by the same enzyme (Gibson et al. 1995). The degradation of halogenated benzoates, arenesulfonates, and nitroarenes, by dioxygenation is discussed in Chapter 9, Parts 1, 3 through 5. Benzoate Dioxygenase In Acinetobacter calcoaceticus, the enzyme is chromosomal and consists of a hydroxylase and an electron transport protein that have been designated benAB and benC (Neidle et al. 1991). The corresponding genes are designated xylXY and xylZ in the plasmid-encoded toluate dioxygenase in Pseudomonas putida, and this dioxygenase accepts a much wider range of substrates.

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Anthranilate-1,2-Dioxygenase In Acinetobacter sp. strain ADP1 (Eby et al. 2001), this enzyme catalyzes the dioxygenation of anthranilate with the loss of both the amino and carboxyl groups. This, like the enzyme from Pseudomonas putida and P. aeruginosa, is a two-component enzyme consisting of an oxygenase and a reductase, whereas the enzyme from Burkholderia cepacia DBO1 consists of three components—a two-subunit oxygenase, a ferredoxin, and a reductase (Chang et al. 2003). Anthranilate may also be hydroxylated to 5-hydroxyanthranilate, and 2,5-dihydroxybenzoate by anthranilate-5-hydroxylase in Nocardia opaca (Cain 1968) and in Ralstonia sp. strain U2 (Zhou et al. 2002), or to 2,3-dihydroxybenzoate in Trichosporon cutaneum (Powlowski et al. 1987) and Aspergillus niger (Subramanian and Vaidyanathan 1984). 2-Halobenzoate 1,2-Dioxygenases These are multicomponent enzymes that catalyze dioxygenation with the loss of the halogen and the carboxyl groups, and are discussed in Chapter 9, Part 1. The application of cis-arene dihydrodiols has been noted in Part 1 of this chapter, and it is sufficient to note here the application of a mutant of Alcaligenes eutrophus strain B9 that is blocked in the degradation of benzoate (and some halogenated benzoates). This produced the cis-1,2-dihydrodiol (Reiner and Hegeman 1971), and has been used as the source of ring B for the synthesis of a range of tetracyclines (Charest et al. 2005). A range of substituted cis-dihydrodiols has been produced, and it has been shown that for 3-substituted benzoates both 3- and 5-substituted cis-dihydrodiols were formed (Reineke et al. 1978).

HYDROXYBENZOATES AND RELATED COMPOUNDS Phenols and hydroxybenzoates are generally degraded by monooxygenation. Illustrative examples include the following: 1. The degradation of salicylate to catechol is initiated by monooxygenation accompanied by decarboxylation (salicylate-1-hydroxylase), and two different and independent salicylate hydroxylases have been found in the naphthalene-degrading Pseudomonas stutzeri AN10 (Bosch et al. 1999). Alternatively, in Rhodococcus sp. strain B4, salicylate is hydroxylated to 2,5-dihydroxybenzoate by salicylate-5-hydroxylase (Grund et al. 1992). An alternative occurs for 5-hydroxy- and 5-aminosalicylate in Pseudaminobacter salicylatoxidans in which ring fission is accomplished directly (Hintner et al. 2001). 2. The monooxygenase from Pseudomonas fluorescens that converts 4-hydroxybenzoate into 3,4-dihydroxybenzoate before ring fission has been characterized (Howell et al. 1972). 3. In Pseudomonas putida, the hydroxylation of 4-hydroxyphenylacetate to 3,4-dihydroxyphenylacetate is carried out by an enzyme that consists of a flavoprotein and a coupling factor (Arunachalam et al. 1992). At least in Escherichia coli, it has been suggested that the large component is an FADH2-utilizing monooxygenase (Xun and Sandvik 2000; Chaiyen et al. 2001). 4. The first enzyme in the degradation of 2-hydroxybiphenyl by Pseudomonas azelaica strain HBP1 is an FAD-dependent monooxygenase that produces 2,3-dihydroxybiphenyl (Suske et al. 1999). Hydroxylation to 1,4-dihydroxy compounds may also activate the ring to oxidative fission. This is illustrated by the following examples of the gentisate pathway: a. The degradation of 3-methylphenol, 3-hydroxybenzoate and salicylate can be initiated by hydroxylation to 2,5-dihydroxybenzoate (gentisate). The range of organisms includes species of Pseudomonas, and Bacillus (Crawford 1975a), and the enteric bacteria Salmonella typhimurium (Goetz and Harmuth 1992) and Klebsiella pneumoniae

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Environmental Degradation and Transformation of Organic Chemicals HO2C

CO2H

CO−CO2H

CO2H

CO2H CO2H CO−CO2H

HO OH

HO2C

OH

OH OH CO2H CO2H

FIGURE 8.26

The gentisate pathway.

(Jones and Cooper 1990). Gentisate dioxygenase carries out fission of the ring to produce pyruvate, and fumarate or maleate (Figure 8.26). Its distribution is noted later. b. The gentisate pathway is used for the degradation of salicylate produced from naphthalene by a Rhodococcus sp. strain B4 (Grund et al. 1992), rather than by the more usual sequence involving the hydroxylative decarboxylation of salicylate to catechol. c. An analogous pathway is used for the degradation of 5-aminosalicylate that is an intermediate in the degradation of 6-aminonaphthalene-2-sulfonate. Direct ring fission of both 5-aminosalicylate and 5-hydroxysalicylate can be accomplished by a salicylate 1,2-dioxygenase in Pseudaminobacter salicylatoxidans (Hintner et al. 2001). d. The gentisate pathway may plausibly be involved in the degradation of benzoate by a denitrifying strain of Pseudomonas sp. in which the initial reaction is the formation of 3-hydroxybenzoate (Altenschmidt et al. 1993). Unusual pathways have been found in the bacterial degradation of a number of 4-hydroxybenzoates and related compounds, and in some of them rearrangements (NIH shifts) are involved: 1. In Pseudomonas putida, l-phenylalanine is hydroxylated to tyrosine by a reaction that involves 6,7-dimethyltetrahydrobiopterin, which is converted into 4a-carbinolamine (Song et al. 1999). Tyrosine is then transformed to 2,5-dihydroxyphenylacetate followed by ring fission (Arias-Barrau et al. 2004). 2. Gentisate is formed by a strain of Bacillus sp. in an unusual rearrangement from 4-hydroxybenzoate (Crawford 1976) that is formally analogous to the formation of 2,5-dihydroxyphenylacetate from 4-hydroxyphenylacetate by Pseudomonas acidovorans (Hareland et al. 1975). Similarly, the metabolism of 4-hydroxybenzoate by the archaeon Haloarcula sp. strain D1 involves the formation of 2,5-dihydroxybenzoate (Fairley et al. 2002). All these reactions putatively involve an NIH shift. 3. The degradation of 4-hydroxyphenylacetate by Pseudomonas acidovorans takes place by hydroxylation with the production of 2,5-dihydroxyphenylacetate in a reaction involving an NIH shift. The final products are fumarate and acetate, and the enzyme has been characterized (Hareland et al. 1975). In contrast, the degradation of 3,4-dihydroxyphenylacetate by the same organism takes place by extradiol fission to produce pyruvate and succinate semialdehyde (Sparnins and Dagley 1975). 4. Pseudomonas sp. strain P.J. 874 grown with tyrosine carried out dioxygenation of 4hydroxyphenylpyruvate to 2,5-dihydroxyphenylacetate accompanied by an NIH shift (Lindstedt et al. 1977). The involvement of a high-spin ferric center coordinated with tyrosine is conclusively revealed in the primary structure of the enzyme (Rüetschi et al. 1992).

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5. The chlorophenol-4-hydroxylase from Burkholderia cepacia strain AC1100 is able to bring about not only the transformation of 4-hydroxybenzaldehydes to the expected 4-hydroxybenzoates, but also rearrangement to 2,5-dihydroxybenzaldehydes (Martin et al. 1999) (Figure 8.27). 6. The degradation of nonylphenol isomers with quaternary side chains in strains of Sphingomonas sp. displays unusual features that involve rearrangement of the side chain (Corvini et al. 2004; Gabriel et al. 2005a,b). This is discussed in Part 4 of this chapter. Both fungi and yeasts are able to degrade simpler substituted aromatic compounds such as vanillate (Ander et al. 1983) (Figure 8.28) and 3,4-dihydroxybenzoate (Anderson and Dagley 1981), both of which involve concomitant decarboxylation and hydroxylation. The oxidative decarboxylation of a range of 4-hydroxybenzoates to 1,4-dihydroxy compounds has been examined, and is carried out by an FAD-dependent monooxygenase in Candida parapsilosis CBS604 (Eppink et al. 1997). The yeast Trichosporon cutaneum is able to metabolize phenol and a number of other aromatic compounds. The pathways by which this is accomplished differ, however, from those that are operative in most bacteria. Illustrative examples include the following (Anderson and Dagley 1980): a. Benzoate is degraded by successive hydroxylation at the 4- and 3-positions followed by concomitant hydroxylation and decarboxylation to benzene-1,2,4-triol, while an analogous pathway is used for 2,5-dihydroxybenzoate. In contrast, 2,3-dihydroxybenzoate is decarboxylated to catechol without hydroxylation. b. The benzene-1,2,4-triol undergoes intradiol ring fission followed by reduction to 3-ketoadipate and incorporation into the TCA cycle.

OH

OH R

R

R

OH R

R

R

or CHO OH

CH O

FIGURE 8.27

CO2H

Anaerobic rearrangement of 4-hydroxybenzaldehydes.

CO2H

OH

OCH3

OH

OCH3

OH

OH

OH

O

O O

O

OH

CO2H CO2H

CO2H CO2H

OH

FIGURE 8.28 Biodegradation of vanillic acid by fungi.

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c. Analogous hydroxylations are used to initiate the metabolism of phenylacetate and 4hydroxyphenyl acetate. Fission of the rings takes place between the ortho hydroxyl group and the –CH2–CO2H group to produce acetoacetate, and fumarate or oxalacetate. Mechanisms for Fission of Oxygenated Rings Catechols In the degradation of the catechols, the next step is the dioxygenation and fission of the ring, either by extradiol (2:3) or intradiol (1:2) ring fission. The intradiol and extradiol enzymes are quite specific for their respective substrates, and whereas all of the first group contain Fe3+, those of the latter contain Fe2+ (Wolgel et al. 1993). Although the extradiol 2,3-dihydroxybiphenyl 1,2-dioxygenases in Rhodococcus globerulus strain P6 are typical in containing Fe (Asturias et al. 1994), the enzyme from Bacillus sp. strain JF8 is manganese-dependent (Hatta et al. 2003). A few other manganesedependent ring-fission dioxygenases have been observed for 3,4-dihydroxyphenylacetate in Arthrobacter globiformis (Boldt et al. 1995) and Bacillus brevis (Que et al. 1981). After formation of the 1,2-dihydroxy compounds, ring fission is mediated by 1:2 (intradiol fission) or 2:3 (extradiol and distal fission) dioxygenases. There are, however, important variations in the pathways used by various groups of microorganisms: 1. The pathways and their regulation during the degradation of catechol and 3,4-dihydroxybenzoate in Pseudomonas putida have been elucidated in extensive studies (Ornston 1966). In this organism, intradiol ring fission is carried out by a 3,4-dioxygenase to produce 3-ketoadipate (Figure 8.29). The stereochemistry of the reactions after ring fission has been examined in detail (Kozarich 1988), and the regulation and genetics in a range of organisms have been reviewed (Harwood and Parales 1996). In contrast, 3,4,5-trihydroxybenzoate (gallate) is degraded in P. putida by extradiol fission with the production of 2 moles of pyruvate (Sparnins and Dagley 1975). 2. In the degradation of 3-hydroxybenzoate, divergent pathways are used by different groups of pseudomonads: a. Pseudomonas testosteroni uses 4,5-dioxygenase to produce pyruvate and formate from 3,4-dihydroxybenzoate by extradiol fission (Figure 8.30a) (Wheelis et al. 1967). The degradation involves pyrone-4,6-dicarboxylate that was converted by a hydrolase into the open-chain 2-hydroxy-4-carboxymuconic acid that was produced directly by extradiol fission from gallate by the same enzyme (Kersten et al. 1982); b. Pseudomonas acidovorans produces 2,5-dihydroxybenzoate that is degraded by gentisate 1,2-dioxygenase to fumarate and pyruvate, which has already been noted (Figure 8.26), and two different gentisate dioxygenases have been characterized (Harpel and Lipscomb 1990). 3. The third alternative for ring fission of 3,4-dihydroxybenzoate is exemplified by Bacillus macerans and B. circulans that use a 2,3-dioxygenase to accomplish this (Figure 8.30b) HO2C

OH

HO2C

CO2H

OH

HO

O

O

O

CO2H

CO2H

CH2CO2H O

CO2H

CH2CO2H CO2H

CO2H +

CH3CO2H

CO2H

FIGURE 8.29 The β-ketoadipate pathway.

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CO2H HO2C

CHO

(a)

CO 2 H

OH

OH

OH CO2H

CO2H

(b) OH OH

CH3−CO−CO2H + H−CO2H

OH

CHO

CHO

CO2H

CO2H

CH3−CO−CO2H + CH3−CHO

OH

FIGURE 8.30 Biodegradation of 3,4-dihydroxybenzoate mediated by (a) 4,5-dioxygenase in Pseudomonas testosteroni and (b) 2,3-dioxygenase in Bacillus macerans.

(Crawford 1975a,b, 1976). In addition, a 2,3-dioxygenase is elaborated by Gram-negative bacteria for the degradation of 3,4-dihydroxyphenylacetate (Sparnins et al. 1974), and by Gram-positive bacteria for the degradation of l-tyrosine via 3,4-dihydroxyphenylacetate (Sparnins and Chapman 1976). 4. The enzymes of alternative pathways may be induced in a given strain by growth with different substrates; for example, growth of Pseudomonas putida R1 with salicylate induces enzymes of the extradiol fission pathway, whereas growth with benzoate induces those of the intradiol pathway (Chakrabarty 1972). As a broad generalization, the extradiol fission is preferred for the degradation of more complex compounds such as toluene, naphthalene, and biphenyl (Furukawa et al. 1983). Methoxybenzoates and Related Compounds Studies on the degradation of lignin have attracted interest in the degradation of the monomeric vanillate and syringate, and mechanisms for their de-O-methylation are given in Chapter 11, Part 2. For substrates such as vanillate and isovanillate, degradation is initiated by de-O-methylation, followed in Comamonas testosteroni by extradiol fission (Providenti et al. 2006). 3-O-methylgallate occupies a central position in several degradations such as that of (a) 4-hydroxy-3-methoxymandelate by Acinetobacter lwoffii (Sze and Dagley 1987), (b) syringate by Sphingomonas paucimobilis SYK-6 (Kasai et al. 2004), and (c) 3,4,5-trimethoxybenzoate by Pseudomonas putida (Donnelly and Dagley 1980). In all of them, dioxygenation carries out extradiol ring fission to an ester that may undergo cyclization to 2-pyrone-4,6-dicarboxylate with the loss of methanol, and degradation of 4-oxalomesaconitate to pyruvate and oxalacetate (Figures 8.31a through 8.31c). 2-Hydroxybenzoate and 1-Hydroxynaphthalene-2-Carboxylate Salicylate is generally degraded by monooxygenation and decarboxylation mediated by salicylate1-hydroxylase to catechol (White-Stevens et al. 1972 a,b), although monooxygenation without decarboxylation can be carried out by salicylate-5-hydroxylase to produce 2,5-dihydroxybenzoate (Zhou et al. 2002). The degradation of naphthalene-2-carboxylate by Burkholderia sp. strain JT 1500 involves the formation of 1-hydroxynaphthalene-2-carboxylate rather than initial oxidative decarboxylation. Naphthalene-1,2-dihydrodiol-2-carboxylate is not, however, involved; hence, the reaction is possibly carried out either by a monooxygenase, or a dehydration step is involved. Subsequent reactions produced pyruvate and o-phthalate, which was degraded via 4,5dihydroxyphthalate (Morawski et al. 1997).

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Environmental Degradation and Transformation of Organic Chemicals CH−CO2H CH(OH)CO2H

OMe

CO2H

MeO

OH

CH

OMe

MeO

OH (a)

OMe MeO

(b)

(c)

CO2H

HO

OMe OH

CO2H

OH

HO2C

CO2Me

HO2C−CO

CO2H

HO2C

O

FIGURE 8.31

CO2H

CO2H

CH3COCO2H

+

HO2C−CO−CH2CO2H

O

Degradation of 3-O-methylgallate.

Fission of 1,4-Dihydroxybenzoates Gentisate (2,5-dihydroxybenzoate) is produced from a range of substrates and has already been noted. Fission of the ring is carried out by gentisate 1,2-dioxygenase that has been characterized from a range of organisms including Moraxella osloensis (Crawford et al. 1975); Pseudomonas testosteroni and P. acidovorans (Harpel and Lipscomb 1990); P. alcaligenes and P. putida (Feng et al. 1999); Sphingomonas sp. strain RW 5 (Wergath et al. 1998); and Klebsiella pneumoniae (Suárez et al. 1996). Nonoxidative Decarboxylation of Benzoate and Related Compounds Although benzoate is generally metabolized by simultaneous dioxygenation and decarboxylation to catechol followed by ring fission, nonoxidative decarboxylation of hydroxybenzoates has been observed. Important issues include the following: (a) the reaction is carried out both by aerobic and anaerobic bacteria, (b) it may be reversible, (c) some of the enzymes—even from aerobes—are oxygen-sensitive, and (d) a high degree of specificity is often observed. 1. Strains of Bacillus megaterium and Streptomyces sp. strain 179 transformed vanillate to guaiacol by decarboxylation (Crawford and Olson 1978), and the gene cluster encoding decarboxylation in Streptomyces sp. strain m D7 has been characterized (Chow et al. 1999). 2. Decarboxylation of aromatic carboxylic acids has been encountered extensively in facultatively anaerobic Enterobacteriaceae. For example, 4-hydroxycinnamic acid is

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CO2H

OMe OH

FIGURE 8.32

OMe OH

Decarboxylation of ferulic acid (3-methoxy-4-hydroxycinnamic acid).

decarboxylated to 4-hydroxystyrene, and ferulic acid (3-methoxy-4-hydroxycinnamic acid) to 4-vinylguaiacol by several strains of Hafnia alvei and H. protea, and by single strains of Enterobacter cloacae and K. aerogenes (Figure 8.32) (Lindsay and Priest 1975). The enzyme has been purified from Bacillus pumilis (Degrassi et al. 1995). 3. The degradation of 2,2′-dihydroxy- 3,3′-dimethoxybiphenyl-5,5′-dicarboxylate (5,5′-dehydrodivanillate) by Sphingomonas paucimobilis SYK-6 proceeds by partial de-O-methylation followed by extradiol fission of the catechol to 2-hydroxy-3-methoxy-5-carboxybenzoate. Diversion of this into central metabolic pathways involves decarboxylation to vanillate by two separate decarboxylases LigW1 and LigW2 (Peng et al. 2005). 4. Decarboxylation is an important step in the degradation of phthalates. The degradation of o- and p-phthalate (terephthalate) has been examined in strains of both Gram-negative and Gram-positive bacteria, and the genetics has been established (references in Sasoh et al. 2006). In summary, degradation is accomplished by dioxygenation followed by dehydrogenation and decarboxylation: a. Dioxygenation initiates the degradation of o-phthalate by Pseudomonas cepacia. A two-component enzyme consisting of a nonheme iron oxygenase and an NADH-dependent oxidoreductase containing FMN and a [2Fe–2S] ferredoxin (Batie et al. 1987) produces the 4,5-dihydrodiol. Degradation is completed by dehydrogenation, decarboxylation (Pujar and Ribbons 1985) to 3,4-dihydroxybenzoate, and ring fission to 2-carboxymuconate. The pathway for the degradation of o-phthalate in Micrococcus sp. strain 12B differs by the initial formation of the 3,4-dihydrodiol followed by dehydrogenation and decarboxylation to 3,4-dihydroxybenzoate (Eaton and Ribbons 1982). The degradation of 4-methyl-o-phthalate by P. fluorescens strain JT701 follows the same pathway by formation of the 2,3-dihydrodiol, followed by decarboxylation with the formation of 4-methyl-2,3-dihydroxybenzoate and extradiol ring fission (Ribbons et al. 1984). b. 3,4-dihydroxybenzoate is produced by an analogous pathway during the degradation of p-phthalate by Comamonas sp. strain T-2 via (Schäfli et al. 1994), and of 5-hydroxyisophthalate via 4,5-dihydroxyisophthalate (Elmorsi and Hopper 1979). The phthalate dioxygenase from Pseudomonas cepacia is also active with pyidine-2,3and 3,4-dicarboxylates (Batie et al. 1987), and this is substantiated in growth experiments in which different strains were able to metabolize pyridine-2,6-, 2,5-, and 2,3-dicarboxylate (Taylor and Amador 1988). It has been shown that there are two regions of DNA that encode the enzymes for o-phthalate degradation, with the gene encoding quinolinate phosphoribosyl transferase located between them, and insertional knockout mutants with elevated levels of this enzyme enhanced growth on o-phthalate (Chang and Zylstra 1999). 5. The reversible decarboxylation of 4-hydroxybenzoate and 3,4-dihydroxybenzoate has been described in Sedimentibacter (Clostridium) hydroxybenzoicum (He and Wiegel 1996), and the oxygen-sensitive enzyme has been purified. The decarboxylase from Pantoae

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agglomerans, which is specific for 3,4,5-trihydroxybenzoate (gallate), is also oxygensensitive (Zeida et al. 1998). 6. The decarboxylase in Clostridium thermoceticum produces CO2 that is essential for growth, and under CO2 limitation, CO2 provides both the methyl and the carboxyl groups of acetate (Hsu et al. 1990). Alternative Pathways for the Degradation of Benzoates and Related Compounds Unusual reactions have been encountered in the aerobic degradations carried out by Azoarcus evansii and Geobacillus stearothermophilus (Zaar et al. 2001). The anaerobic degradation of benzoate by Azoarcus evansii (Ebenau-Jehle et al. 2003) and Thauera aromatica (Dörner and Boll 2002), and of 3-hydroxybenzoate by Th. aromatica (Laempe et al. 2001) is discussed later. 1. a. Aerobic degradation of benzoate in Azoarcus evansii takes place via the CoAester to produce ∆1,2-adipate and 3-ketoadipate, that is degraded to CoA-esters of succinate and acetate (Zaar et al. 2001; Gescher et al. 2002) (Figure 8.33). The initial dihydrodiol has been characterized using 13C NMR, and the enzyme that carries this out consists of two components: BoxA that is an Fe-S protein and functions as a reductase, and BoxB that functions as an oxygenase (Zaar et al. 2004). Fission of the dihydrodiol is mediated by a lyase BoxC that is not dependent on oxygen and results in the production of formate from C-2 (Gescher et al. 2005). The product from dehydrogenation by BoxD is then converted into 3-ketoadipate by conventional steps (Gescher et al. 2006). b. Burkholderia xenovorans strain LB400 uses several pathways for the degradation of the benzoate produced by degradation of biphenyl. One of them is the classical benzoate dihydroxylation/catechol intradiol fission pathway, while another resembles the pathway that is used by Azoarcus evansii. The genes for this pathway occur both in the chromosome when they are expressed during growth with biphenyl, and in a megaplasmid copy where they were detected only in benzoate-grown cells in the transition to the stationary phase (Denef et al. 2005). 2. Novel pathways for the aerobic degradation of anthranilate (2-aminobenzoate) were described several years ago in a strain then designated as a Pseudomonas sp. (Lochmeyer et al. 1992; Altenschmidt and Fuchs 1992), and now reassigned to Azoarcus evansii. The pathway is analogous to that used for benzoate by this strain. This is initiated by formation of the benzoyl SCoA-ester followed by monooxygenation/ reduction to 2-amino-5-ketocyclohex-1-ene that undergoes β-oxidation with a concomitant NIH shift of the hydrogen at C5 to C6 (Hartmann et al. 1999; Schühle et al. 2001) (Figure 8.34).

CO−SCoA

CO−SCoA

CO−SCoA

CO−SCoA

OH CHO

OH CO−SCoA

CO2H

CO−SCoA O

CO2H

FIGURE 8.33

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CO2H

Alternative aerobic degradation of benzoyl-CoA by Azoarcus evansii.

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Carbocyclic Aromatic Compounds without Halogen Substituents CoS−SCoA

R N

N H

N

H

NH2 H

N

HO

O O OH

Co−SCoA H

NH2 +

NADH

O

NH2

H

H

O H

FIGURE 8.34

CoS−CoA

NH2

O

Co−SCoA H

433

H

H

Aerobic degradation of 2-aminobenzoate by Azoarcus evansii.

AEROBIC REDUCTION OF ARENE CARBOXYLATES The reduction of aromatic carboxylic acids to the corresponding aldehydes under aerobic conditions is of interest in biotechnology, since the oxidoreductase from Nocardia sp. is able to accept a range of substituted benzoic acids, naphthoic acids, and a few heterocyclic carboxylic acids (Li and Rosazza 1997). The reaction involves formation of an acyl-AMP intermediate by reaction of the carboxylic acid with ATP; NADPH then reduces this to the aldehyde (Li and Rosazza 1998; He et al. 2004). A comparable reaction for aromatic carboxylates has been demonstrated in Neurospora crassa (Gross 1972).

ARENES WITH AN OXYGENATED C2 OR C3 SIDE CHAIN Several pathways are used for the aerobic degradation of aromatic compounds with an oxygenated C2 or C3 side chain. These include acetophenones and reduced compounds that may be oxidized to acetophenones, and compounds including tropic acid, styrene, and phenylethylamine that can be metabolized to phenylacetate, which has already been discussed. 1. The mandelate pathway in Pseudomonas putida involves successive oxidation to benzoyl formate and benzoate, which is further metabolized via catechol and the 3-ketoadipate pathway (Figure 8.35a) (Hegeman 1966). Both enantiomers of mandelate were degraded through the activity of a mandelate racemase (Hegeman 1966), and the racemase (mdlA) is encoded in an operon that includes the next two enzymes in the pathway—S-mandelate dehydrogenase (mdlB) and benzoylformate decarboxylase (mdlC) (Tsou et al. 1990). A formally comparable pathway is used by a strain of Alcaligenes sp. that degrades 4-hydroxyacetophenone to 4-hydroxybenzoyl methanol, which is oxidized in an unusual reaction to 4-hydroxybenzoate and formate. The 4-hydroxybenzoate is then metabolized to 3-ketoadipate via 3,4-dihydroxybenzoate (Figure 8.35b) (Hopper et al. 1985). 2. Although the metabolism of vanillate generally involves de-O-methylation to 3,4-dihydroxybenzoate followed by intradiol ring fission, in Acinetobacter lwoffii vanillate is hydroxylated to 3-O-methyl gallate, which produces pyruvate and oxalacetate in reactions that have already been noted (Sze and Dagley 1987). 3. The metabolism of ferulate to vanillin by Pseudomonas fluorescens strain AN103 is carried out by an enoyl-SCoA hydratase/isomerase rather than by oxidation, and the enzyme belongs to the enoyl-CoA hydratase superfamily (Gasson et al. 1998).

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CO2H

CH(OH)

CO

CO2H

OH OH

(a)

CH3

CH2OH

CO

CO

CO2H

CO2H

OH

OH

OH

OH

(b) OH

CH3 CO

O−COCH3

OH

(c)

FIGURE 8.35 Degradation of (a) mandelate, (b) 4-hydroxyacetophenone by side-chain oxidation pathways, (c) acetophenone by Baeyer–Villiger monooxygenation.

4. A strain of Pseudomonas sp. AT3 degraded tropic acid, which has a –CH–(CH2OH)– CO2H side chain to phenylacetate (Long et al. 1997). 5. Baeyer–Villiger-type oxidations initiate the degradation of acetophenone by strains of Arthrobacter sp. and Nocardia sp. (Cripps et al. 1978), and of 4-hydroxyacetophenone by Pseudomonas putida strain JD1 (Darby et al. 1987). Acetophenone is converted into phenyl acetate, which is hydrolyzed to phenol and then hydroxylated to catechol before ring fission (Figure 8.35c). Similarly, 4-hydroxyacetophenone is oxidized to 4hydroxyphenyl acetate, which is hydrolyzed to 1,4-dihydroxybenzene before ring fission to 3-ketoadipate. The 4-hydroxyacetophenone monooxygenase from Pseudomonas fluorescens has been purified from recombinant Escherichia coli, and was able to carry out Baeyer–Villiger oxidation of a range of substituted acetophenones, bicyclic cyclobutanones, and the enantiomerically specific oxidation of methyl aryl sulfides to sulfoxides (Kamerbeek et al. 2003). The metabolism of chloroacetophenones takes place by analogous monooxygenation to the corresponding chlorophenylesters. Some ortho-substituted chlorophenols are, however, inhibitory, and only low rates of oxidation have been encountered with di- and trichlorinated acetophenones, so that growth with them is not possible (Higson and Focht 1990). However, degradation of 4-chloroacetophenone has been demonstrated with a mixed culture of an Arthrobacter sp. and a Micrococcus sp. (Havel and Reineke 1993). 6. Phenylalanine-ammonia lyase (PAL), which catalyzes the nonoxidative deamination of l-phenylalanine to trans-cinnamate, is widespread among plants as a precursor of lignins, flavanoids, and coumarins but is uncommon among bacteria. The gene has been characterized in “Streptomyces maritimus” (Xiang and Moore 2002), and the homologous l-histidine ammonia lyase that produces trans-urocanate from l-histidine is the first enzyme in the pathway for the bacterial degradation of l-histidine.

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CH2

COSCoA COSCoA OH OH

435

COSCoA

COSCoA CH3COSCoA +

O

FIGURE 8.36

CO2H

CO2H

Aerobic degradation of phenylacetate.

7. Some unusual reactions have emerged in the degradation of phenylacetate and 3-phenyl butyrates: a. The degradation of hydroxylated phenylacetates and related compounds such as tyrosine has been described. These include 3,4-dihydroxyphenylacetate by Escherichia coli (Cooper and Skinner 1980), and 2,5-dihydroxyphenylacetate produced by rearrangement of 4-hydroxyphenylacetate in Xanthobacter sp. strain 124X (van der Tweel et al. 1986). However, these are not intermediates in the degradation of phenylacetate, and an important alternative for this has been described. The degradation of phenylacetate by Escherichia coli is carried out by a pathway, which has been elucidated in detail using 13C NMR of the labeled substrate (Ismail et al. 2003). It is initiated by the formation of the coenzyme A-ester, and continues with the formation of a 1,2-diol, which undergoes reduction and fission of the ring (Figure 8.36) in reactions that are formally analogous to the oxygenation/reduction in the degradation of benzoate by Azoarcus evansii. This pathway is also followed by Escherichia coli strain W (Fernández et al. 2006) and Rhodococcus sp. strain RHA1, which has an established ability to degrade PCBs and a wide range of monocyclic aromatic compounds (Navarro-Llorens et al. 2005). b. Although the aerobic degradation of 3-phenylpropionate by Escherichia coli proceeded as expected by ring dioxygenation, dehydrogenation, and ring fission to pyruvate, succinate, and acetate (Díaz et al. 1998), the degradation of iboprofen (2-[4-isobutylphenyl]-propionic acid) with a −CH(CH3)CO2H side chain was different. The catechol produced by dioxygenation and dehydrogenation was succeeded exceptionally by loss of the side chain (Murdoch and Hay 2005), involving a hitherto unresolved mechanism. c. i. The degradation of 3-phenylbutyrate was examined in a strain of Pseudomonas sp. (Sariaslani et al. 1982). The formation of the expected catechol was proved by NMR analysis of its nonenzymatic transformation to a dihydrocoumarin, and characterization of the ring fission product by reaction with NH3 to produce a pyridine. In addition, however, products that were identified by mass spectrometry (MS) included 3-phenylpropionate and 3-phenylacetate. It was tentatively suggested that the first of these was produced by oxidation of the methyl group followed by decarboxylation, although the expected phenylsuccinate could not be utilized. ii. The degradation of (R)- and (S)-3-phenyl butyrate has been examined in Rhodococcus rhodochrous strain PB1. Whereas the (S)-enantiomer was dioxygenated to the 2,3dihydroxy compound that was not metabolized further, the (R)-enantiomer was converted into 3-phenylpropionate by a mechanism that was not established. This was followed by dioxygenation and fission of the ring to succinate (Simoni et al. 1996). Anaerobic Metabolism Considerable effort has been devoted to the anaerobic degradation of aromatic compounds. It is important to note that several distinct groups of organisms are involved: (a) strictly anaerobic

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fermentative bacteria, (b) strictly anaerobic photoheterotrophic bacteria, (c) anaerobic sulfate-reducing bacteria, and (d) organisms using nitrate as electron acceptor under anaerobic conditions. Pragmatically, these organisms generally belong to the following groups: methanogenic, sulfidogenic, denitrifying, and phototrophic. Although the greatest attention is given here to studies using pure cultures and investigations in which the relevant enzymes have been characterized, some valuable examples from studies using mixed cultures are provided. Bacteria Using Nitrate Electron Acceptor under Anaerobic Conditions, and Anaerobic Phototrophs It has become clear that benzoate occupies a central position in the anaerobic degradation of both phenols and alkylated arenes such as toluene and xylenes, and that carboxylation, hydroxylation, and reductive dehydroxylation are important reactions for phenols that are discussed in Part 4 of this chapter. The simplest examples include alkylated benzenes, products from the carboxylation of napthalene and phenanthrene (Zhang and Young 1997), the decarboxylation of o-, m-, and p-phthalate under denitrifying conditions (Nozawa and Maruyama 1988), and the metabolism of phenols and anilines by carboxylation. Further illustrative examples include the following: 1. Under denitrifying conditions, 2-aminobenzoate is degraded by a Pseudomonas sp. to benzoate, which is then reduced to cyclohexene-1-carboxylate (Lochmeyer et al. 1992). 2. The metabolism of cinnamate and ω-phenylalkane carboxylates has been studied in Rhodopseudomonas palustris (Elder et al. 1992), and for growth with the higher homologs additional CO2 was necessary. The key degradative reaction was β-oxidation, for compounds with chain lengths of three, five, and seven carbon atoms, benzoate was formed and further metabolized, but for the even-numbered compounds with four, six, and eight carbon atoms phenylacetate was a terminal metabolite. 3. The anaerobic metabolism of l-phenylalanine by Thauera aromatica under denitrifying conditions involves several steps that result in the formation of benzoyl-CoA: (a) conversion to the CoA-ester by a ligase, (b) transamination to phenylacetyl-CoA, (c) α-oxidation to phenylglyoxalate, and (d) decarboxylation to benzoyl-CoA (Schneider et al. 1997). An analogous pathway is used by Azoarcus evansii (Hirsch et al. 1998). The membrane-bound phenylacetyl-CoA: acceptor oxidoreductase that is induced under denitrifying conditions during growth with phenylalanine or phenylacetate has been purified (Schneider and Fuchs 1998). The level of the enzyme was low in cells grown with phenylglyoxalate, the enzyme was insensitive to oxygen, and was absent in cells grown aerobically with phenylacetate. 4. Phenylacetate and 4-hydroxyphenylacetate are oxidized sequentially under anaerobic conditions by a denitrifying strain of Pseudomonas sp. to the phenylglyoxylates and benzoate (Mohamed et al. 1993). Benzoate The pathway for the anaerobic degradation of benzoate is entirely different from that used under aerobic conditions, and in contrast to degradation under aerobic conditions (with the exception of those carried out by Azoarcus evansii in which this is used under aerobic conditions), it invariably involves initial formation of the CoA-esters. The coenzyme A-ester is an inducer for both the anaerobic and aerobic pathways depending on the absence or presence of oxygen (Schüle et al. 2003). The pathway for its anaerobic degradation under denitrifying or phototrophic conditions involves three cardinal reactions: (1) formation of benzoyl-CoA, (2) partial reduction of the ring, and (3) ring fission (references in Harwood et al. 1999) (Figures 8.37a and 8.37b). The designation of the enzymes corresponds to those for the genes discussed below: 1. Formation of the coenzyme A thioester is mediated by a ligase, which has been demonstrated in a number of organisms including Pseudomonas sp. strain K172

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437

SCoA C

O

SCoA C

(b)

(a) O

SCoA

Dch

O

SCoA C

C HO

BadK

Had O

O

SCoA

SCoA C

C

OH

O

BadH O

O

SCoA

O

SCoA C

C OH

O

BadI

Oah

CO−SCoA

CO2H CO−SCoA HO

FIGURE 8.37

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CO2H

Anaerobic degradation of benzoate.

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2.

3. 4.

5.

6.

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(Thauera sp. strain K172) (Dangel et al. 1991) and Rhodopseudomonas palustris (Egland et al. 1995). A benzoyl-CoA reductase reduces one or more of the double bonds in the aromatic ring (Gibson and Gibson 1992; Koch and Fuchs 1992). In Thauera aromatica, the reduction is mediated by a 2-ketoglutarate: ferredoxin oxidoreductase (Dörner and Boll 2002), whereas in Azoarcus evansii, a 2-ketoglutarate, NADP+ oxidoreductase is involved (Ebenau-Jehle et al. 2003). This has been suggested as a biomarker for the anaerobic degradation of aromatic compounds by developing degenerate primers from the established genes of Thauera aromatica and Azoarcus evansii. The isolated clones were then assigned to the bcr or bzd type of the reductase (Song and Ward 2005). The cyclohexa-1,5-dienecarboxylate thioester is hydrated to produce the 6-hydroxy compound (Boll et al. 2000). The pathways in T. aromatica and Rh. palustris differ slightly in details. a. In T. aromatica strain K172, the gene sequence for the following enzymes has been determined: bcrCBAD for the benzoyl-CoA reductase, dch for the dienyl-CoA hydratase, and had for 6-hydroxycyclohex-1-ene-1-carboxy-CoA dehydrogenase, and they occur in the order had, dch, bcrCBAD (Breese et al. 1998). b. In Rh. palustris, the genes are badDEFG for the benzoyl-CoA reductase, badK for the cyclohex-1ene-1-carboxyl-CoA hydratase, badH for the 2-hydroxycyclohexane1-carboxyl-CoA dehydrogenase, and badL for the ketohydrolase. They occur in the order badK, badL, badH, badD, badE, badF, badG, the first three transcribing in the direction opposite to the others (Harwood et al. 1999; Egland and Harwood 1999). Fission of the ring: In T. aromatica, the initial ring fission product is hepta-3-ene-1,7dicarboxylyl-CoA formed by 3-ketoacyl-CoA hydrolase, which is encoded by oah that lies immediately to the right of had. The resulting 3-hydroxpimelyl-CoA is degraded by β-oxidation to glutaryl-CoA. With the minor modifications noted above, the same pathway has been demonstrated in Rhodopseudomonas palustris, and the ring fission enzyme 2-ketocyclohexanecarboxyl-CoA hydrolase that produces pimelyl-CoA has been purified (Pelletier and Harwood 1998). The gene for the ring-fission enzyme badL lies to the left of badH. It has been shown in Rh. palustris that the suit of enzymes in the sequence from cyclohexen1-carboxylate onward was induced during growth on benzoate, though not on succinate (Perrotta and Harwood 1994). The further degradation of glutarate takes place through a pathway involving dehydrogenation and decarboxylation to crotonyl-CoA and subsequent formation of acetate (Härtel et al. 1993). After the growth of a denitrifying organism with either benzoate or pimelate, both glutaryl-CoA dehydrogenase and glutaconyl-CoA decarboxylase activities were induced—though not the enzymes leading from pimelyl-CoA to glutaryl-CoA. This is consistent with the involvement of 3-hydroxypimelyl-CoA in the degradation of benzoyl-CoA (Gallus and Schink 1994). The initial steps in the degradation of glutarate involving dehydrogenation, decarboxylation, and the β-oxidation pathway have also been demonstrated in other organisms that produce butyrate and isobutyrate (Matties and Schink 1992). Details of the various pathways for the anaerobic metabolism of acetate have been reviewed (Thauer et al. 1989).

Hydroxybenzoates There are different pathways for the degradation of hydroxybenzoates: 1. For hydroxybenzoates with hydroxyl groups at the ortho or meta positions, degradation is initiated by decarboxylation. a. For 2,6- and 3,5-dihydroxybenzoates, a denitrifying organism initiates degradation by decarboxylation to 1,3-dihydroxybenzene followed by the reduction of the ring to cylohexan-1,3-dione (Kluge et al. 1990);

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b. For 3,4,5- and 2,4,6-trihydroxybenzoate, Pelobacter acidigalli carries out decarboxylation to 1,3,5-trihydroxybenzene followed by reduction of the ring (Brune and Schink 1992). 2. The degradation of 4-hydroxybenzoate by Thauera aromatica (Brackmann and Fuchs 1993) and by the anaerobic phototroph Rhodopseudomonas palustris (Gibson et al. 1997) is initiated by formation of the CoA-ester. The next step involves dehydroxylation to benzoate by a reductase that been characterized in both Thauera aromatica (Breese and Fuchs 1998) and Rh. palustris (Gibson et al. 1997). The benzoate that is produced then enters the established pathway for reduction of the ring (Heider et al. 1998). The degradation of 3-hydroxybenzoate by Sporotomaculum hydroxybenzoicum uses an analogous pathway (Müller and Schink 2000). 3. The fermentation of 3-hydroxybenzoate by Sporotomaculum hydroxybenzoicum produces acetate, butyrate, and CO2, with benzoate as a transient intermediate (Brauman et al. 1998). However, although the degradation of 3-hydroxybenzoate by Thauera aromatica begins with the formation of the CoA-ester, this is followed by the reduction of the ring with retention of the original hydroxyl group (Laempe et al. 2001).

ALDEHYDES The reductive transformation of arene carboxylates to the corresponding aldehydes under aerobic conditions has already been noted. In addition, aromatic aldehydes may undergo both reductive and oxidative reactions, with the possibility of decarboxylation of the carboxylic acid formed: a. Oxidation of substituted benzaldehydes to benzoates at the expense of sulfate reduction has been demonstrated in strains of Desulfovibrio sp., although the carboxylic acids produced were apparently stable to further degradation (Zellner et al. 1990). Vanillin was, however, used as a substrate for growth by a strain of Desulfotomaculum sp. and was metabolized via vanillate and catechol (Kuever et al. 1993). Vanillin is transformed by Clostridium formicoaceticum sequentially to vanillate and 3,4-dihydroxybenzoate, while the methyl group is converted into acetate via acetyl coenzyme A (Göbner et al. 1994). b. Reduction of benzoates to the corresponding benzyl alcohols has been observed in several organisms: (i) cell extracts of Clostridium formicoaceticum at the expense of carbon monoxide (Fraisse and Simon 1988), (ii) Desulfomicrobium escambiense at the expense of pyruvate that was oxidized to acetate, lactate, and succinate (Sharak Genther et al. 1997). Oxidoreductases mediate the transfer of electrons between aldehydes and carboxylates. In Clostridium formicoaceticum, there are two oxidoreductases. One contains Mo, is reversible, and active toward a range of both aliphatic and aromatic carboxylic acids and aldehydes (White et al. 1993), whereas the other is a W-containing enzyme (White et al. 1991). In addition, there exists a different enzyme that carries out the same reaction in C. thermoaceticum (White et al. 1989). Different aldehyde oxidoreductases have been isolated from Clostridium thermoaceticum and Cl. formicoaceticum (White et al. 1993). Whereas however, both cinnamate and cinnamaldehyde were good substrates for the Mo-containing enzyme from the latter, benzoate was an extremely poor substrate. The W-containing enzyme from Cl. formicoaceticum, however, displays high activity toward a greater range of substituted benzoates (White et al. 1991).

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Crawford RL (1976) Pathways of 4-hydroxybenzoate degradation among species of Bacillus. J Bacteriol 127: 204–290. Crawford RL, PP Olson (1978) Microbial catabolism of vanillate: decarboxylation to guaiacol. Appl Environ Microbiol 36: 539–543. Crawford RL, SW Hutton, PJ Chapman (1975) Purification and properties of gentisate 1,2-dioxygenase from Moraxella osloensis. J Bacteriol 121: 794–799. Cripps RE, PW Trudgill, JG Whateley (1978) The metabolism of 1-phenylethanol and acetophenone by Nocardia T5 and an Arthrobacter species. Eur J Biochem 86: 175–186. Dangel W, R Brackmann, A Lack, M Mohamed, J Koch, B Oswald, B Seyfried, A Tschech, G Fuchs (1991) Differential expression of enzyme activities initiating anoxic metabolism of various aromatic compounds via benzoyl-CoA. Arch Microbiol 155: 256–262. Darby JM, DG Taylor, DJ Hopper (1987) Hydroquinone as the ring-fission substrate in the catabolism of 4-ethylphenol and 4-hydroxyacetophenone by Pseudomonas putida D1. J Gen Microbiol 133: 2137–2146. Degrassi G, PP de Laureto, CV Bruschi (1995) Purification and chracterization of ferulate and p-coumarate decarboxylase from Bacillus pumilis. Appl Environ Microbiol 61: 326–332. Denef VJ, MA Patrauchan, C Florizone, J Park, TV Tsoi, W Verstraete, JM Tiedje, LD Eltis (2005) Growth substrate- and phase-specific expression of biphenyl, benzoate, and C1 metabolic pathways in Burkholderia xenovorans LB400. J Bacteriol 187: 7996–8005. Díaz E, A Ferrández, JL Garcia (1998) Characterization of the hca cluster encoding the dioxygenolytic pathway for initial catabolism of 3-phenylpropionic acid in Escherichia coli K-12. J Bacteriol 180: 2915–2923. Donnelly MI, S Dagley (1980) Production of methanol from aromatic acids by Pseudomonas putida. J Bacteriol 142: 916–924. Dörner E, M Boll (2002) Properties of 2-oxoglutarate: ferredoxin oxidoreductase from Thauera aromatica and its role in enzymatic reduction of the aromatic ring. J Bacteriol 184: 3975–3983. Eaton RW, DW Ribbons (1982) Metabolism of dibutylphthalate and phthalate by Micrococcus sp. strain 12B. J Bacteriol 151: 48–57. Ebenau-Jehle C, M Boll, G Fuchs (2003) 2-oxoglutarate: NADP+ oxidoreductase in Azoarcus evansii: properties and function in electron transfer reactions in aromatic ring reduction. J Bacteriol 185: 6119–6129. Eby DM, ZM Beharry, ED Coulter, DM Kurtz, EL Neidle (2001) Characterization and evolution of anthranilate 1,2-dioxygenase from Acinetobacter sp. strain ADP1. J Bacteriol 183: 109–118. Egland PG, CS Harwood (1999) BadR, a new MarR family member regulates anaerobic benzoate degradation by Rhodopseudomonas palustris in concert with AadR, an Fnr family member. J. Bacteriol. 181: 2102–2109. Egland PG, J Gibson, CS Harwood (1995) Benzoate-coenzyme A ligase, encoded badA, is one of three lighases to catalyze benzoyl-coenzyme A formation during anaerobic growth of Rhodopseudomonas palustris on benzoate. J Bacteriol 177: 6545–6551. Elder DJE, P Morgan, DJ Kelly (1992) Anaerobic degradation of trans-cinnamate and ω-phenylalkane carboxylic acids by the photosynthetic bacterium Rhodopseudomonas palustris: evidence for a betaoxidation mechanism. Arch Microbiol 157: 148–154. Elmorsi EA, DJ Hopper (1979) The catabolism of 5-hydroxyisophthalate by a soil bacterium. J Gen Microbiol 111: 145–152. Eppink MHM, SA Boeren, J Vervoort, WJH van Berkel (1997) Purification and properties of 4-hydroxybenzoate 1-hydroxylase (decarboxylating), a novel flavin adenine dinucleotide-dependent monooxygenase from Candida parapilosis CBS604. J Bacteriol 179: 6680–6687. Fairley DJ, DR Boyd, ND Sharma, CCR Allen, P Morgan, MJ Larkin (2002) Aerobic metabolism of 4hydroxybenzoic acid in Archaea via an unusual pathway involving an intramolecular migration (NIH shift). Appl Environ Microbiol 68: 6246–6255. Feng Y, HE Khoo, CL Poh (1999) Purification and characterization of gentisate 1,2-dioxygenases from Pseudomonas alcaligenes NCIB 9867 and Pseudomonas putida NCIB 9869. Appl Environ Microbiol 65: 946–950. Fernández C, A Ferrández, B Minambres, E Diaz, JL Gardia (2006) Genetic characterization of the phenylacetyl-coenzyme A oxygenase from the aerobic phenylacetic acid degradation pathway of Escherichia coli. Appl Environ Microbiol 72: 7422–7426.

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PART 4 NONHALOGENATED PHENOLS AND ANILINES Although phenols and anilines have similar electronic structures, the initial step in their aerobic degradation is different. Whereas the degradation of phenols is initiated by monooxygenation, dioxygenation with loss of the amino group is used for anilines. Catechols are then produced from both of them. The degradation of halogenated phenols and anilines is discussed in Chapter 9, Part 2.

PHENOLS Aerobic Degradation The degradation of phenol is initiated by hydroxylation using a single-component NAD(P)H-dependent monooxygenase encoded by pheA (Nurk et al. 1991). In Bacillus thermoglucosidasius, however, the hydroxylase is a two-component system encoded by pheA1 and pheA2, and hydroxylation requires supplementation by the second component with FAD(NADH) (Kirchner et al. 2003). For alkylated phenols, there are different pathways for degradation depending on whether or not oxidation of the alkyl group precedes oxygenation and fission of the ring: a. The degradation of cresols and xylenols by strains of Pseudomonas putida is initiated by oxidation of one of the methyl groups to carboxylate before ring hydroxylation and ring

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fission (Bayly et al. 1988). For example, 4-methyl phenol is oxidized to 4-hydroxybenzoate, which is hydroxylated to 3,4-dihydroxybenoate and undergoes ring fission to 3-ketoadipate. For 3-methylphenol, degradation of the 3-hydroxybenzoate involves hydroxylation to 2,5dihydroxybenzoate (gentisate), which is degraded to pyruvate and maleate (Hopper and Chapman 1971). b. Pseudomonas sp. strain KL 28 is able to degrade 4-n-alkylphenols with chain lengths of one to five carbon atoms. Degradation of both the 3-n-alkyl and 4-n-alkylphenols is initiated by hydroxylation to 3,4-dihydroxy compounds that are degraded by extradiol fission followed by dehydrogenation to 2-hydroxy-5-n-alkylmuconates with the alkyl groups intact (Jeong et al. 2003). Some unusual reactions have been encountered in the degradation of substituted phenols: 1. The degradation of bisphenol-A involves a rearrangement—analogous to those described for compounds with quaternary carbon atoms in Chapter 7, Part 1—followed by oxidative fission of the stilbene produced (Lobos et al. 1992; Spivack et al. 1994) (Figure 8.38). 2. The degradation of 4-ethylphenol and related compounds is initiated not by oxygenation but by dehydrogenation to a quinone methide followed by hydroxylation (Hopper and Cottrell 2003) (Figure 8.39), and the flavocytochrome 4-ethylphenol methylene hydroxylase in Pseudomonas putida stain JD1 has been characterized (Reeve et al. 1989). CH3

CH3

CH3

HO

OH

CH2OH

HO

OH

OH

OH CH2 OH

HO

CH3

HO

OH

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CO2H

COCH2OH

OH

OH

CH3

HO

FIGURE 8.38

CH2OH

HO

CO2H

COCH3

OH

OH

Degradation of bisphenol-A.

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CH3

CH2

OH

O

OH

OH

FIGURE 8.39

Aerobic degradation of 4-ethylphenol. OH

OH

H C

C OH

O

O

OH

(a)

O C

OH

C

O

(b)

OH

O

O C

HO

FIGURE 8.40

OH O

O

O

C

OH + O

C

Biotransformations of nonylphenol.

3. The degradation of nonylphenol isomers with quaternary side chains in a strain of Sphingomonas xenophaga displays unusual features (Corvini et al. 2004; Gabriel et al. 2005a,b). These may be described formally as oxygenation at the position para to the phenol group, followed by either of two rearrangements: a. Migration of the alkyl group to C-3 of the phenolic ring with the formation of an alkylated hydroquinone (Figure 8.40a); b. Rearrangement by insertion of an oxygen atom from the hydroxyl group between the terminal position of the alkyl group and C-4 of the phenolic ring, followed by further oxygenation with the loss of the tertiary alkanol and formation of benzoquinone (Figure 8.40b). Analogous reactions may be presumed to have taken place in the transformation of a tert-octylphenol by a strain of Sphingomonas sp. in which 2,4,4-trimethylpent-2-anol was produced (Tanghe et al. 2000). The degradation of polyhydroxybenzenes has been examined rather seldom. 1. The degradation of resorcinol can take place by several pathways that are initiated by hydroxylation. 1,2,4-Trihydroxybenzene is degraded by (a) 1:2 dioxygenation and

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ring fission to maleylacetate and 3-ketoadipate or 1:6 dioxygenation and ring fission to acetylpyruvate in strains of Pseudomonas putida (Chapman and Ribbons 1976), or (b) 1:2 dioxygenation and ring fission in Corynebacterium glutamicum (Huang et al. 2006). In contrast, degradation by Azotobacter vinelandii involved hydroxylation to pyrogallol that was degraded by intradiol fission to 2-hydroxymuconate and pyruvate (Groseclose and Ribbons 1981). 2. The metabolism of phloroglucinol (1,3,5-trihydroxybenzene) by Fusarium solani produced 1,2,3-trihydroxybenzene (pyrogallol) as an unexpected intermediate, which was degraded by intradiol fission to vinylpyruvate, and then to acetaldehyde and pyruvate (Walker and Taylor 1983). Under anaerobic conditions, Pelobacter massiliensis uses the transhydroxylation pathway for the degradation of all the trihydroxbenzenes (Brune et al.1992). 3. 2,4-Diacetylphloroglucinol, which is a secondary metabolite produced by Pseudomonas fluorescens CHA0, is the effective biocontrol agent for fungal pathogens. As a 1:3-diketone, it can be hydrolyzed to the monoacetyl compound, which is much less toxic to Fusarium oxysporum, and finally to phloroglucinol (Schouten et al. 2004). The hydrolase has been characterized from Pseudomonas fluorescens CHA0, is pecific for 2,4-diacetylphloroglucinol, and was unreactive against 2-acetyl- and 2,4,6-triacetylphloroglucinol (Bottiglieri and Keel 2006). Anaerobic Degradation The anaerobic degradation of phenols is dependent on two reactions—carboxylation of the ring to hydroxybenzoates and dehydroxylation to benzoates. Ring carboxylation is a key reaction in the anaerobic degradation of both phenols and anilines. In Thauera aromatica, the carboxylation of phenol is carried out in two steps: AMP-dependent phosphorylation by a synthase followed by carboxylation with the production of 4-hydroxybenzoate (Schühle and Fuchs 2004; Narmandakh et al. 2006). Catechol and aniline can be degraded to CO2 by Desulfobacterium anilini (Schnell et al. 1989), and catechol is slowly degraded by D. catecholicum (Szewzyk and Pfennig 1987), and by a strain of Desulfotomaculum sp. (Kuever et al. 1993). Analogy with phenol and consistent with the CO2 dependence of the catechol-degrading Desulfotomaculum sp. (Kuever et al. 1993), catechol is carboxylated to 3,4-dihydroxybenzoate by a strain of Desulfobacterium sp. (Gorny and Schink 1994b). This is then dehydroxylated to benzoate before further degradation. The reversible decarboxylation of 3,4-dihydroxybenzoate has been described in Sedimentibacter (Clostridium) hydroxybenzoicum (He and Wiegel 1996), and the enzyme has been purified. 4-HydroxybenzoylCoA plays a central role in the metabolism of phenol, 4-methylphenol, 4-hydroxyphenylacetate, and related aromatic compounds by a denitrifying strain of Pseudomonas sp. (now classified as Thauera aromatica). This is supported by the isolation and purification of the dehydroxylating enzyme (Brackmann and Fuchs 1993), and the characterization of the enzyme as a molybdenumflavin-iron-sulfur protein (Breese and Fuchs 1998). A similar dehydroxylase is also involved in the metabolism of 4-hydroxybenzoate by Rhodopseudomonas palustris (Gibson et al. 1997), and is also a molybdenum-containing enzyme. The degradation of 3-methylphenol by oxidation to 3-hydroxybenzoate followed by analogous dehydroxylation has been observed in a “Pseudomonas-like” organism under denitrifying conditions (Bonting et al. 1995), and in Desulfotomaculum sp. strain Gross (Londry et al. 1997). The oxidation is formally comparable to the reactions in the metabolism of alkylarenes, and a carboxylation pathway is apparently not involved. The degradation of phenol is initiated by formation of phenylphosphate followed by carboxylation and dehydroxylation to benzoate. Benzoyl CoA is then degraded to acetate (Breinig et al. 2000) or—under methanogenic conditions to methane. An analogous pathway to benzoyl CoA has been demonstrated for hydroquinone (Gorny and Schink 1994a). In contrast, the degradation of 1,3-dihydroxybenzene (resorcinol) by Azoarcus anaerobius under denitrifying conditions apparently proceeds by hydroxylation to 1,2,4-trihydroxybenzene followed by oxidation at the expense

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of nitrate to 2-hydroxybenzoquinone. Putative enzymes for established ring-fission reactions or hydrolysis could not be found, and the reactions leading to fission of the aromatic ring remain to be elucidated (Phillip and Schink 1998). The initial reactions involved in the degradation of 2-, 3-, or 4-methylphenol (o-, m-, p-cresol) are quite different. Degradation in Desulfitobacterium cetonicum is initiated by reaction with fumarate catalyzed by benzylsuccinate synthase—analogous to that already noted for the anaerobic degradation of toluene and xylenes (Verfürth et al. 2004). For m- and p-cresol, this is followed by dehydrogenation to 3- or 4-hydroxybenzoyl-CoA and dehydroxylation to benzoyl-CoA (Figure 8.41) (Müller et al. 2001). The degradation of 3-methylphenol (m-cresol) may alternatively proceed by oxidation to 3-hydroxybenzoate followed by dehydroxylation to benzoate (Londry et al. 1997). Gallate (3,4,5-trihydroxybenzoate) can be degraded by Pelobacter acidigallici to acetate and CO2 (Schink and Pfennig 1982), and by Eubacterium oxidoreducens in the presence of exogenous H2 or formate to acetate, butyrate, and CO2 (Krumholz et al. 1987). The degradation has been studied in detail, and takes place by this unusual pathway (Figure 8.42) (Brune and Schink 1992; Krumholtz and Bryant 1988). The formation of phloroglucinol in P. acidigallici involves a series of intramolecular hydroxyl transfer reactions with regeneration of 1,2,3,5-tetrahydroxybenzene (Brune and Schink 1990) (Figure 8.43), and in Eubacterium oxidoreducens the reaction is catalyzed by an isomerase that contains Fe, acid-labile sulfur, and Mo (Krumholtz and Bryant 1988). Although

CoSCoA

CoSCoA

CO2H

CO2H

CoSCoA

CH2

CH3

OH

OH

FIGURE 8.41

OH

OH

OH

Anaerobic degradation of 4-methylphenol.

HO

OH

OH

HO

OH

HO

O

OH HO

CH3−CO−CH2−CH(OH)−CH2−CO2H

CH3−CO−CH2−CO−CH2−CO2H

3CH3−CO2H

Pathway for the biodegradation of 3,4,5-trihydroxybenzoate. OH

OH

OH

+ HO

OH OH

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O

OH

OH

FIGURE 8.43

CoSCoA

CO2H

CO2H

FIGURE 8.42

CoSCoA

CO

+ HO

OH OH

HO

OH

HO

OH

OH

Intramolecular hydroxyl group transfers in the biodegradation of 1,2,3-trihydroxybenzoate.

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the reaction is a transhydroxylase, it does not involve hydroxylation by water, which is found in many other molybdenum enzymes (Reichenbecher et al. 1996). An analogous reaction occurs in E. oxidoreducens and also involves 1,2,3,5-tetrahydroxybenzene, although details of the pathway may be different (Haddock and Ferry 1989, 1993). Subsequent reduction to dihydrophloroglucinol is followed by ring fission to 3-hydroxy-5-ketohexanoate and, finally, to formation of butryrate and acetate (Krumholz et al. 1987). This pathway is comparable to that involved in the degradation of resorcinol by the species of Clostridium (Tschech and Schink 1985), although an alternative pathway involving direct hydrolysis has been observed in denitrifying bacteria (Gorny et al. 1992). P. massiliensis is able to degrade all three trihydroxybenzenes (1,2,3-, 1,3,5-, and 1,2,4-) by the transhydroxylation pathway (Brune et al. 1992), whereas Desulfovibrio inopinatus degrades 1,2,4trihydroxybenzene by a different pathway (Reichenbecher and Schink 1997). The degradation of phenols has also been examined in mixed cultures, and embodies some of the features that have been observed in pure cultures, in particular, the significance of carboxylation. The following examples illustrate some additional details of these reactions. It is worth noting that metabolism of the methylphenols (cresols) by mixed cultures may proceed by ring carboxylation in contrast to the oxidation pathway that has been demonstrated in pure cultures: 1. Phenol is carboxylated by a defined obligate syntrophic consortium to benzoate, which is then degraded to acetate, methane, and CO2 (Knoll and Winter 1989). 2. 2-Methylphenol is carboxylated by a methanogenic consortium to 4-hydroxy-3-methylbenzoate that was dehydroxylated to 3-methylbenzoate, which was the stable end product (Bisaillon et al. 1991). 3. 3-Methylphenol is carboxylated to 2-methyl-4-hydroxybenzoate by a methanogenic-enrichment culture before degradation to acetate (Figure 8.44a) (Roberts et al. 1990): 14C-labeled bicarbonate produced carboxyl-labeled acetate, while 14C-methyl-labeled 3-methylphenol yielded methyl-labeled acetate. However, 2-methylbenzoate formed by dehydroxylation of 2-methyl-4-hydroxybenzoate was not further metabolized (Figure 8.44b). A similar reaction occurs with a sulfate-reducing mixed culture (Ramanand and Suflita 1991). 4. An unusual reaction occurred during the degradation of 3-methylphenol by a methanogenic consortium (Londry and Fedorak 1993). Although carboxylation to 2-methyl-4hydroxybenzoate took place as in the preceding example, further metabolism involved loss of the methyl group with the formation of methane before dehydroxylation to benzoate (Figure 8.44c).

CH3−CO2H OH

(a)

OH

(b) CH3

CH3

CH3

(c)

CO2H

CO2H OH

CO2H

FIGURE 8.44

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CO2H

Anaerobic biodegradation of 3-methylphenol.

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It can be concluded from these observations that whereas benzoate produced by the carboxylation of phenols can be degraded, dehydroxylation with the formation of substituted benzoates may produce stable terminal metabolites. Anilines Aerobic The aerobic degradation of anilines is, in principle, straightforward and involves deamination in a strain of Nocardia sp. by a dioxygenase (Bachofer and Lingens 1975), although details of the enzyme have not been fully resolved (Fukumori and Saint 1997). This is followed by ring fission of the resulting catechols by either intradiol or extradiol ring fission. Analogous mechanisms are used for anthranilic acid by deamination and decarboxylation (Chang et al. 2003), and 2-aminobenzenesulfonate to catechol-3-sulfonate with concomitant deamination (Junker et al. 1994). The degradation of 3- and 4-methylanilines by Pseudomonas putida mt-2 has been described (McClure and Venables 1986, 1987). It is important to emphasize that since anilines may be incorporated into humic material, their fate is not determined solely by biodegradation. The aerobic degradation of aminobenzenesulfonates that are produced from azobenzenes by reduction and scission is discussed in Chapter 9, Part 4. Anaerobic In analogy with phenol, aniline is carboxylated to 4-aminobenzoate followed by reductive deamination to benzoate (Schnell and Schink 1991), which is degraded by pathways that have already been elaborated.

REFERENCES Bachofer R, F Lingens (1975) Conversion of aniline into pyrocatechol by a Nocardia sp. : incorporation of oxygen-18. FEBS Lett 50: 288–290. Bayly R, R Jain, CL Poh, R Skurry (1988) Unity and diversity in the degradation of xylenols by Pseudomonas spp.: a model for the study of microbial evolution. In Microbial Metabolism and the Carbon Cycle (Eds SR Hagedorn, RS Hanson, DA Kunz) pp. 359–379. Harwood Academic Publishers, Chur, Switzerland. Bisaillon J-G, F Lépine, R Beaudet, M Sylvestre (1991) Carboxylation of o-cresol by an anaerobic consortium under methanogenic conditions. Appl Environ Microbiol 57: 2131–2134. Bonting CFC, S Schneider, G Schmidtberg, G Fuchs (1995) Anaerobic degradation of m-cresol via methyl oxidation to 3-hydroxybenzoate by a denitrifying bacterium. Arch Microbiol 164: 63–69. Bottiglieri M, C Keel (2006) Characterization of PhlG, a hydrolase that specifically degrades the antifungal compound diacetylphloroglucinol in the biocontrol agent Pseudomonas fluorescens CHA0. Appl Environ Microbiol 72: 418–427. Brackmann R, G Fuchs (1993) Enzymes of anaerobic metabolism of phenolic compounds 4-hydroxy benzoyl-CoA reductase (dehydroxylating) from a denitrifying Pseudomonas sp. Eur J Biochem 213: 563–571. Breese K, G Fuchs (1998) 4-hydroxybenzoyl-CoA reductase (dehydroxylating) from the denitrifying bacterium Thauera aromatica: prosthetic groups, electrom donor, and genes of a member of the molybdenumflavin-iron-sulfur proteins. Eur J Biochem 251: 916–923. Breinig S, E Schiltz, G Fuchs (2000) Genes involved in anaerobic metabolism of phenol in the bacterium Thauera aromatica. J Bacteriol 182: 5849–5863. Brune A, B Schink (1990) Pyrogallol-to-phloroglucinal conversion and other hydroxyl-transfer reactions catalyzed by cell extracts of Pelobacter acidigallici. J Bacteriol 172: 1070–1076. Brune A, B Schink (1992) Phloroglucinol pathway in the strictly anaerobic Pelobacter acidigallici fermentation of trihydroxybenzenes to acetate via triacetic acid. Arch Microbiol 157: 417–424. Brune A, S Schnell, B Schink (1992) Sequential transhydroxylations converting hydroxyhydroquinone to phloroglucinol in the strictly anaerobic, fermentative bacterium. Pelobacter massiliensis. Appl Environ Microbiol 58: 1861–1868.

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Chang H-K, P Mohseni, GJ Zylstra (2003) Characterization and regulation of the genes for a novel anthanilate 1,2 dioxygenase from Burkholderia cepacia DBO 1. J Bacteriol 185: 5871–5881. Chapman PJ, DW Ribbons (1976) metabolism of resorcinylic compounds by bacteria: alternative pathways fore resorcinol catabolism in Pseudomonas putida. J Bacteriol 125: 985–998. Corvini PFX, RJW Meesters, A Schäffer, HF Schröder, R Vinken, J Hollender (2004) Degradation of a nonylphenol single isomer by Sphingomonas sp. strain TTNP3 leads to a hydroxylation-induced migration product. Appl Environ Microbiol 70: 6897–6900. Fukumori F, CP Saint (1997) Nucleotide sequences and regulational analysis of genes involved in conversion of aniline to catechol in Pseudomonas putida UCC22 (pTDN1). J Bacteriol 179: 399–408. Gabriel FLP, A Heidlberger, D Rentsch, W Giger, K Guenther, H-P E Kohler (2005b) A novel metabolic pathway for degradation of 4-nonylphenol environmental contaminants by Sphingomonas xenophaga Bayram. J Biol Chem 280: 15526–15533. Gabriel FLP, W Giger, K Guenther, H-P E Kohler (2005a) Differential degradation of nonylphenol isomers by Sphingomonas xenophaga Bayram. Appl Environ Microbiol 71: 1123–1129. Gibson J, M Dispensa, CS Harwood (1997) 4-hydroxybenzoyl coenzyme A reductase dehydroxylating is required for anaerobic degradation of 4-hydrozybenzoate by Rhodopseudomonas palustris and shares features with molybdenum-containing hydroxylases. J Bacteriol 179: 634–642. Gorny N, B Schink (1994a) Hydroquinone degradation via reductive dehydroxylation of gentisyl-CoA by a strictly anaerobic fermenting bacterium. Arch Microbiol 161: 25–32. Gorny N, B Schink (1994b) Anaerobic degradation of catechol by Desulfobacterium sp strain cat2 proceeds via carboxylation to protocatechuate. Appl Environ Microbiol 60: 3396–3400. Gorny N, G Wahl, A Brune, B Schink (1992) A strictly anaerobic nitrate-reducing bacterium growing with resorcinol and other aromatic compounds. Arch Microbiol 158: 48–53. Groseclose EE, DW Ribbons (1981) Metabolism of resorcinolic compounds by bacteria: a new pathway for resorcinol catabolism in Azotobacter vinelandii. J Bacteriol 146: 460–466. Haddock JD, JG Ferry (1989) Purification and properties of phloroglucinol reductase from Eubacterium oxidoreducens G-41. J Biol Chem 264: 4423–4427. Haddock JD, JG Ferry (1993) Initial steps in the anaerobic degradation of 3,4,5-trihydroxybenzoate by Eubacterium oxidoreducens: characterization of mutants and role of 1,2,3,5-tetrahydroxybenzene. J Bacteriol 175: 669–673. He Z, J Wiegel (1996) Purification and characterization of an oxygen-sensitive, reversible 3,4-dihydroxybenzoate decarboxylase from Clostridium hydroxybenzoicum. J Bacteriol 178: 3539–3543. Hopper DJ, L Cottrell (2003) Alkylphenol biotransformations catalyzed by 4-ethylphenol methylenehydroxylase. Appl Environ Microbiol 69: 3650–3652. Hopper DJ, PJ Chapman (1971) Gentisic acid and its 3- and 4-methyl-substituted homologues as intermediates in the bacterial degradation of m-cresol, 3,5-xylenol and 2,4-xylenol. Biochem J 122: 19–28. Huang Y, K-X Zhao, X-H Shen, MT Chaudhry, C-Y Jiang, S-J Liu (2006) Genetic characterization of the resorcinol catabolic pathway in Corynebacterium glutamicum. Appl Environ Microbiol 72: 7238–7245. Jeong JJ, JH Kim, C-K Kim, I Hwang, K Lee (2003) 3- and 4-alkylphenol degradation pathway in Pseudomonas sp. strain KL28: genetic organization of the lap gene cluster and substrate specificities of phenol hydroxylase and catechol 2,3-dioxygenase. Microbiology (UK) 149: 3265–3277. Junker F, T Leisinger, AM Cook (1994) 3-sulfocatechol 2,3-dioxygenase and other dioxygenases (EC 1.13.11.2 and EC 1-14-12) in the degradative pathways of 2-aminobenzenesulfonic, benzenesulphonic and 4-toluenesulfonic acids in Alcaligenes sp. strain O-1. Microbiology U.K. 140: 1713–1722. Kirchner U, AH Westphal, R Müller, WJH van Berkel (2003) Phenol hydroxylase from Bacillus thermoglucosidasius A7, a two-component monooxygenase with a dual role for FAD. J Biol Chem 278: 47545–47553. Knoll G, J. Winter (1989) Degradation of phenol via carboxylation to benzoate by a defined, obligate syntrophic consortium of anaerobic bacteria. Appl. Microbiol. Biotechnol. 30: 318–324. Krumholz LR, MP Bryant (1988) Characterization of the pyrogallol–phloroglucinol isomerase of Eubacterium oxidoreducens. J Bacteriol 170: 2472–2479. Krumholz LR, RL Crawford, ME Hemling, MP Bryant (1987) Metabolism of gallate and phloroglucinol in Eubacterium oxidoreducens via 3-hydroxy-5-oxohexanoate. J Bacteriol 169: 1886–1890. Kuever J, J Kulmer, S Janssen, U Fischer, K-H Blotevogel (1993) Isolation and characterization of a new spore-forming sulfate-reducing bacterium growing by complete oxidation of catechol. Arch Microbiol 159: 282–288. Lobos JH, TK Leib, T-M Su (1992) Biodegradation of bisphenol A and other bisphenols by a gram-negative aerobic bacterium. Appl Environ Microbiol 58: 1823–1831.

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Londry KL, PM Fedorak (1993) Use of fluorinated compounds to detect aromatic metabolites from m-cresol in a methanogenic consortium: evidence for a demethylation reaction. Appl Environ Microbiol 59: 2229–2238. Londry KL, PM Fedorak, JM Suflita (1997) Anaerobic degradation of m-cresol by a sulfate-reducing bacterium. Appl Environ Microbiol 63: 3170–3175. McClure NC, WA Venables (1986) Adaptation of Pseudomonas putida mt-2 to growth on aromatic amines. J Gen Microbiol 132: 2209–2218. McClure NC, WA Venables (1987) pTDN1, a catabolic plasmid involved in aromatic amine catabolism in Pseudomonas putida mt-2. J Gen Microbiol 133: 2073–2077. Müller JA, AS Galuschko, A Kappler, B Schink (2001) Initiation of anaerobic degradation of p-cresol by formation of 4-hydroxybenzylsuccinate in Desulfitobacterium cetonicum. J Bacteriol 183: 752–757. Narmandakh A, N Gad′on, F Drepper, B Knapp, W Haehnel, G Fuchs (2006) Phosphorylation of phenol by phenylphosphate synthase: role of histidine phosphate in catalysis. J Bacteriol 188: 7815–7822. Phillip B, B Schink (1998) Evidence of two oxidative reaction steps initiating anaerobic degradation of resorcinol (1,3-dihydroxybenzene) by the denitrifying bacterium Azoarcus anaerobius. J Bacteriol. 180: 3644–3649. Ramanand K, JM Suflita (1991) Anaerobic degradation of m-cresol in anoxic aquifer slurries: carboxylation reactions in a sulfate-reducing bacterial enrichment. Appl. Environ. Microbiol. 57: 1689–1695. Reeve CD, MA Carver, DJ Hopper (1989) The purification and characterization of 4-ethylphenol methylene hydroxylase, a flavocytochrome from Pseudomonas putida JD1. Biochem J 263: 431–437. Reichenbecher W, A Rüdiger, PMH Kroneck, B Schink (1996) One molecule of molybdopterin guanine dinucleotide is associated with each subunit of the heterodimeric Mo-Fe-S protein transhydroxylase of Pelobacter acidigallici as determined by SDS (PAGE) and mass spectrometry. Eur J Biochem 237: 406–413. Reichenbecher WW, B Schink (1997) Desulfovibrio inopinatus sp nov, a new sulfate-reducing bacterium that degrades hydroxyhydroquinone (1,2,4-trihydroxybenzene). Arch Microbiol 168: 338–344. Roberts DJ, PM Fedorak, SE Hrudey (1990) CO2 incorporation and 4-hydroxy-2-methylbenzoic acid formation during anaerobic metabolism of m-cresol by a methanogenic consortium. Appl Environ Microbiol 56: 472–478. Schink B, N Pfennig (1982) Fermentation of trihydroxybenzenes by Pelobacter acidigallici gen nov sp nov, a new strictly anaerobic, non-sporeforming bacterium. Arch Microbiol 133: 195–201. Schnell S, B Schink (1991) Anaerobic aniline degradation via reductive deamination of 4-aminobenzoyl-CoA in Desulfobacterium anilini. Arch Microbiol 155: 183–190. Schnell S, F Bak, N Pfennig (1989) Anaerobic degradation of aniline and dihydroxybenzenes by newly isolated sulfate-reducing bacteria and description of Desulfobacterium anilini. Arch Microbiol 152: 556–563. Schouten A, G van den Berg, C Edel-Hermann, C Steinberg, N Gautheron, C Alabouvette, CH de Vos, P Lemanceau, JM Raaijmakers (2004) Defense responses of Fusarium oxysporum to 2,4-diacetylphloroglucinol, a broad-spectrum antibiotic produced by Pseudomonas fluorescens. Mol Plant–Microbe Interact 17: 1201–1211. Schühle K, G Fuchs (2004) Phenylphosphate carboxylase: a new C–C lyase involved in anaerobic phenol metabolism in Thauera aromatica. J Bacteriol 186: 4556–4567. Spivack J, TK Leib, JH Lobos (1994) Novel pathway for bacterial metabolism of bisphenol A. Rearrangements and stilbene cleavage in bisphenol A metabolism. J Biol Chem 269: 7323–7329. Szewzyk R, N Pfennig (1987) Complete oxidation of catechol by the strictly anaerobic sulfate-reducing Desulfobacterium catecholicum sp. nov. Arch Microbiol 147: 163–168. Tanghe T, W Dhooge, W Verstraete (2000) Formation of the metabolic intermediate 2,4,5-trimethyl2-pentanol during incubation of a Sphingomonas sp. strain with the xeno-estrogenic octylphenol. Biodegradation 11: 11–19. Tschech A, B Schink (1985) Fermentative degradation of resorcinol and resorcylic acids. Arch Microbiol 143: 52–59. Verfürth K, AJ Pierik, C Leutwein, S Zorn, J Heider (2004) Substrate specificities and electron paramagnetic resonance properties of benzylsuccinate synthesis in anaerobic toluene and m-xylene metabolism. Arch Microbiol 181: 155–162. Walker JRL, BG Taylor (1983) Metabolism of phloroglucinol by Fusarium solani. Arch Microbiol 134: 123–126.

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Substituted Carbocyclic Aromatic Compounds

PART 1 HALOGENATED ARENES AND CARBOXYLATES WITH CHLORINE, BROMINE, OR IODINE SUBSTITUENTS These industrial chemicals are widely used as the precursors for a wide range of products, which include agrochemicals, pharmaceutical products, polychlorinated biphenyls (PCBs), and polybrominated diphenyl ether flame retardants. For almost all of them, serious concern has arisen over their adverse environmental effects.

INTRODUCTION Since the aerobic degradation of halogenated phenols takes place by monooxygenation and is discussed in Part 2 of this chapter, it is not discussed here except to note the production of chlorocatechols from chlorophenols and chloroanilines. Emphasis is placed on chlorinated substrates, and reference may be made to a review (Allard and Neilson 2003) for details of their brominated and iodinated analogs. The degradation of aromatic fluorinated compounds is discussed in Part 3 of this chapter. Aerobic degradation of chlorinated arene hydrocarbons, including the important group PCBs, and chlorobenzoates that are produced from them as metabolites, is generally initiated by dihydroxylation of the rings to dihydrodiols followed by dehydrogenation to catechols. Halide may be lost simultaneously and for 2-halogenated benzoates, both halide and carboxyl. Salient aspects are summarized, and attention drawn to selected aspects of enzyme inhibition. Halogenated catechols play a central role in the degradation of halogenated aromatic hydrocarbons, carboxylates, phenols, and amines. Fission of 3-substituted catechols may take place by extradiol, intradiol, or distal fission (Figure 9.1). To avoid the production of toxic metabolites, the halogenated catechols generally use both intradiol (1:2) and—exceptionally—distal (1:6) fission, whereas the alkylated catechols generally involve extradiol fission (2:3) (Figure 9.2). Further details are given later in this section.

HALOGENATED ARENE HYDROCARBONS Aerobic Conditions 1. Initial dioxygenation. Although degradation generally proceeds by dioxygenation followed by dehydrogenation of the cis-dihydrodiols that are formed, dehydrogenation may be replaced by elimination to produce catechols directly, for example, in the degradation of 1,2,4,5-tetrachlorobenzene (Sander et al. 1991), and chlorinated biphenyls with chlorine at the 2-positions when 2:3 dioxygenation has taken place (Seeger et al. 2001). Analogously, 2-halogenated benzoates generally undergo dioxygenation with concomitant decarboxylation and loss of halide, whereas other isomers are degraded by dioxygenation followed by dehydrogenation to chlorocatechols. 2. Ring fission. Ring fission of substituted catechols may take place by three pathways: intradiol (ortho), extradiol (meta), or distal (1:6) fission. Although extradiol fission is generally preferred for substituted catechols, the product from extradiol fission of 3-chlorocatechol is normally inhibitory since the acyl chloride reacts with the enzyme 455

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Cl

(a)

(c)

(b) OH CO 2 H CHO

OH

CO 2 H CO 2 H

Cl

Cl

CO 2 H COCl

FIGURE 9.1 Alternative dioxygenation pathways for 3-substituted catechols (a) distal (b) intradiol (c) extradiol.

R CO 2 H R

CHO

1,2

OH R OH CO 2 H CHO

FIGURE 9.2

2,3

Ring-cleavage pathways for the biodegradation of 3-substituted catechols.

producing it. Degradation then takes place by what has been termed the “modified ortho pathway” with the production of chloromuconates. As an alternative, spontaneous hydrolysis of the acyl chloride is able to relieve the inhibition so that the extradiol pathway is available, for example, in Pseudomonas putida strain GJ31 (Mars et al. 1997; Kaschabek et al. 1998). An analogous example is found in the degradation of 2-chlorohydroquinone (Endo et al. 2005). Further examples are given in Chapter 3, Part 1. 3. Additional limitations. Other limitations have been encountered: (a) the formation of the inhibitory protoanemonin from 4-chlorocatechols (Blasco et al. 1995) and (b) the ineffectiveness of cycloisomerases required for effective degradation of 2-chlorotoluene (Pollmann et al. 2005). An example of this limitation is given later, and in the degradation of 3-fluorobenzoate in Part 3 of this chapter.

MONOCYCLIC CHLORINATED ARENES Ultimately, ring cleavage of halogenated aromatic hydrocarbons must occur if aerobic degradation—rather than merely biotransformation—is to be accomplished. The pathways outlined for unsubstituted hydrocarbons generally followed are formation of a cis-dihydrodiol followed by dehydrogenation to the catechol and ring cleavage to muconic acids or hydroxymuconate semialdehydes. Dioxygenation, dehydrogenation, and intradiol fission of the resulting catechols are used in the degradation of 1,2,4-trichlorobenzene (Figure 9.3) and 1,2,3,4-tetrachlorobenzene by Pseudomonas chlororaphis strain RW71 (Potrawfke et al. 1998). A strain of Pseudomonas sp. PS12 can degrade 1,2,4-trichlorobenzene by a comparable pathway, although chloride is lost before the production of 3,4,6-trichlorocatechol during the degradation of 1,2,4,5-tetrachlorobenzene (Figure 9.4) (Sander et al. 1991). The genes encoding the α- and β-subunits of the terminal oxidase, the ferredoxin, and

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Substituted Carbocyclic Aromatic Compounds Cl

Cl

Cl

Cl OH

Cl

457

Cl

Cl OH

Cl

CO 2 H

H H Cl

Cl

Cl

Cl

O

CO 2 H

O

CO 2 H

Cl

Cl

Biodegradation of 1,2,4-trichlorobenzene.

Cl

Cl

Cl OH

Cl

Cl

Cl

OH

H Cl Cl Cl

FIGURE 9.4

HO 2 C

O

Cl

FIGURE 9.3

Cl

Cl HO 2 C

HO 2 C O

CO 2 H

OH

OH

OH

OH Cl

Cl

Biodegradation of 1,2,4,5-tetrachlorobenzene.

the reductase from this strain now designated Burkholderia sp. have been cloned into Escherichia coli (Beil et al. 1997). It was confirmed that whereas metabolism of 1,2- and 1,4-dichlorobenzene and 1,2,4-tri- and 1,2,3,4-tetrachlorobenzene produced stable cis-dihydrodiols, dechlorination concomitant with dioxygenation took place with 1,2,4,5-tetrachlorobenzene, and was confirmed using 18 O2. In addition, dioxygenation of biphenyl, dibenzo [1,4]dioxin, dibenzofuran, and naphthalene was observed in this constructed strain. An additional example of simultaneous dechlorination and dioxygenation in the degradation of PCB congeners is given later. The importance of both the number and the position of chlorine substituents is illustrated by the results of a study with Xanthobacter flavus strain 14p1. The strain grew only with 1,4-dichlorobenzene, which was degraded by dioxygenation, dehydrogenation, and intradiol ring fission, whereas 1,3-dichlorobenzene did not induce dioxygenase activity (Sommer and Görisch 1997). Although chlorobenzene induced several of the enzymes essential for degradation, its inability to support growth was assumed to be the result of the accumulation of toxic intermediates. The transformation of chlorotoluenes has been studied in Ralstonia (Pseudomonas) sp. strain PS12, and important reasons that limit their biodegradability have emerged (Pollmann et al. 2001). a. Whereas 2,4-, 2,5-, and 3,4-dichlorotoluenes were dioxygenated and dehydrogenated to the corresponding catechols—and were therefore potentially degradable, 2,3-, 2,6-, and 3,5-dichlorotoluenes and 2,4,5-trichlorotoluene were only monooxygenated at the methyl group to the benzyl alcohols. b. Although the 2,3-, 2,4-, 2,5-, and 2,6-dichlorobenzyl alcohols could be oxidized to the benzoate, these apparently could not be degraded for unresolved reasons. c. 2-Chlorotoluene appears to be recalcitrant in this strain, although dioxygenation and dehydrogenation to chloromethylcatechols took place: the cycloisomerases were unable to degrade effectively the muconolactones produced by intradiol ring fission (Pollmann et al. 2005).

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Environmental Degradation and Transformation of Organic Chemicals CCl3 CH

Cl

CCl 3 Cl

OH

FIGURE 9.5

Cl

CH

OH

OH

Alternative degradation of DDT.

Although 1,1-dichloro-2,2-bis(p-chlorophenyl)ethene (DDE) is appreciably recalcitrant, it can be degraded by dioxygenation and extradiol ring fission to chlorobenzoate by cells of Pseudomonas acidovorans M3GY grown with biphenyl (Hay and Focht 1998; Aislabie et al. 1999). Bis (p-chlorophenyl)acetic acid (DDA), which is a polar metabolite, is apparently persistent in the environment (Heberer and Dünnbier 1999). Alternatively, degradation of DDT may take place by hydroxylation of the ring and displacement of the aromatic ring chlorine atom by hydroxyl (Figure 9.5) (Massé et al. 1989). Mutations at the active site of CYP101 (cytochrome P450cam) from a strain of Pseudomonas putida made possible the monooxygenation of chlorinated benzenes with less than three substituents to chlorophenols, with concomitant NIH shifts for 1,3-dichlorobenzene (Jones et al. 2001). Further mutations made it possible to oxidize even pentachlorobenzene and hexachlorobenzene to pentachlorophenol (Chen et al. 2002). Integration of the genes encoding cytochrome P450cam into Sphingobium chlorophenolicum enabled this strain to partially transform hexachlorobenzene to pentachlorophenol (Yan et al. 2006). Anaerobic Conditions Attention has been directed to the dechlorination of polychlorinated benzenes by strains that use them as an energy source by dehalorespiration. Investigations using Dahalococcoides sp. strain CBDB1 have shown its ability to dechlorinate congeners with three or more chlorine substituents (Hölscher et al. 2003). Although there are minor pathways, the major one for hexachlorobenzene was successive reductive dechlorination to pentachlorobenzene, 1,2,4,5-tetrachlorobenzene, 1,2,4trichlorobenzene, and 1,4-dichlorobenzene (Jayachandran et al. 2003). The electron transport system has been examined by the use of specific inhibitors. Ionophores had no effect on dechlorination, whereas the ATP-synthase inhibitor N,N′-dicyclohexylcarbodiimide (DCCD) was strongly inhibitory (Jayachandran et al. 2004).

POLYCHLORINATED BIPHENYLS Aerobic Degradation This is discussed in some detail both on account of the importance of the degradation of PCBs, which are widely distributed contaminants, and since it illustrates a number of important principles. Overview The degradation of PCBs has attracted enormous attention. Although impressive investigation has been directed to Burkholderia xenovorans strain LB400 on account of the wide spectrum of PCBs that are oxidized, there are a number of factors that determine PCB biodegradability. These include (a) both the number and position of the chlorine substituents, (b) the positions at which dioxygenation occurs, (c) whether or not the initially formed dihydrodiol can be dehydrogenated, and (d) the degradability of the metabolites produced by ring fission. Considerable effort has been made to enlarge the range of strains that are susceptible to degradation by genetic manipulation. Because PCBs are lipophilic contaminants, a range of biota including fish, fish-eating birds, and mammals can accumulate PCBs. In contrast to the reactions described in this section, these are metabolized by the formation of sulfones in mammals (Letcher et al. 1998) or by monooxygenation with the

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production of hydroxylated PCBs in fish (Buckman et al. 2006). The anaerobic dechlorination of these metabolites has been described (Wiegel et al. 1999) in the context of the degradation of halogenated phenols. In this section, the strains used have been given to minimize confusion because there have been a number of changes in the nomenclature of some organisms, for example, Acinetobacter sp. strain P6 → Rhodococcus globerulus strain P6 Pseudomonas sp. strain LB400 → Burkholderia xenovorans strain LB400 Pseudomonas pseudoalcaligenes KF707 → Pseudomonas paucimobilis KF707 The assignments by the original authors have been used, with the alternatives in parentheses. A range of organisms able to degrade PCB congeners has been isolated, generally from biphenyl enrichments. These include a. Gram-negative strains—Alcaligenes eutrophus strain H850 (Bedard et al. 1987a), Pseudomonas pseudoalcaligenes strain KF707 (Furukawa and Miyazaki 1986), Pseudomonas (Burkholderia) sp. strain LB400 (Bopp 1986), and P. testosteroni (Ahmad et al. 1990) b. Gram-positive strains—Rhodococcus globerulus strain P6 (Acinetobacter sp. strain P6) (Furukawa et al. 1979; Asturias and Timmis 1993), Rhodococcus erythropolis strain TA421 (Maeda et al. 1995), Rhodococcus sp. strain RHA1 (Seto et al. 1995), and Bacillus sp. strain JF8 (Shimura et al. 1999) There are considerable differences in the oxidative activity of these organisms toward PCB congeners. For example, (a) Burkholderia (Pseudomonas) sp. strain LB400 has a much greater versatility than P. pseudoalcaligenes strain KF707 (Gibson et al. 1993) or P. testosteroni (Ahmad et al. 1990) and (b) the range of congeners transformed by Rhodococcus sp. strain RHA1 differs from that of Burkholderia sp. strain LB400 and Pseudomonas paucimobilis strain KF707 (Seto et al. 1995). Degradation of PCBs is carried out by a suite of enzymes comprising biphenyl-2,3-dioxygenase, biphenyl-2,3-dihydrodiol dehydrogenase, 2,3-dihydroxybiphenyl dioxygenase, and the hydrolytic enzymes that produce benzoate encoded by the genes bphA, bphB, bphC, and bphD (Furukawa and Miyazaki 1986; Ahmad et al. 1991; Taira et al. 1992). Degradation Is Initiated by Dioxygenation Dioxygenation generally takes place at the 2,3-position, or less commonly, at the 3,4-position, to produce cis-2,3-dihydrodiols or cis-3,4-dihydrodiols. For example, a strain of Pseudomonas sp. that was able to degrade a range of PCB congeners had both 2,3-dioxygenase and 3,4-dioxygenase activity, and four of the open reading frames were homologous to components of toluene dioxygenase (Erickson and Mondelo 1992). It was not resolved whether a single dioxygenase was able to introduce oxygen at the 2,3- or the 3,4-positions, or whether there were two different enzymes. The dioxygenase has been purified and characterized from several organisms including Burkholderia (Pseudomonas) sp. strain LB400 (Haddock and Gibson 1995) and Comomonas testosteroni strain B-356 (Hurtubise et al. 1995). Dioxygenation required three components—a [2Fe–2S] oxygenase ISPBDH, a [2Fe–2S] ferredoxin (FERBPH), and a flavin reductase (FERBPH). These are encoded by bphA (α-subunit), bphE (β-subunit), bphF, and bphG. The iron–sulfur protein consists of a large subunit (α) with a molecular mass of 51–52 kDa and a small subunit (β) with a molecular mass of 22–27 kDa. Whole-cell studies using oxygen uptake by Burkholderia (Pseudomonas) sp. strain LB400 had clearly shown the versatility of this organism in the degradation of PCB congeners containing up

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to four chlorine substituents. The biphenyl 2,3-dioxygenase from this strain has been purified and revealed additional mechanistic details (Haddock et al. 1995): a. The dihydrodiol was produced from all congeners, and as an alternative to dehydrogenation, an ortho-chlorine substituent could be eliminated from the 2,3-dihydrodiol to produce the catechol. b. 4,5-Dihydrodiols were produced from 3-3 (3,3′), (25-2) 2,2′,5, and the (25-25) 2,2′,5,5′ congeners, although these are not substrates for the dihydrodiol dehydrogenase. These results are consistent with the results accumulated over many years, and with the most frequent operation of a single dioxygenase enzyme. Several issues determine the specificity of strains for the degradation of PCBs, and extensive efforts have been made to increase the range of strains that can be degraded. The specificity reflects differences in the protein sequences in two regions (III and IV) of the BphA large α-subunit of the dioxygenase (Kimura et al. 1997; Mondello et al. 1997), and this offers the possibility of expanding the range of degradable PCBs by alterations of specific amino acids. The purified dioxygenase from Burkholderia sp. strain LB400—whicht is one of the most versatile—has been examined in the presence of electron-transport proteins and cofactors for its reactivity toward congeners (Arnett et al. 2000). Reactivity depended both on the number and position of substituents, and some congeners were recalcitrant, including (24-4) 2,4,4′-trichloro-, (24-24) 2,4,2′,4′-tetrachloro-, and (345-25) 3,4,5,2′,5′-pentachlorobiphenyl. Substantial effort has therefore been directed to elucidating determinants of the specificity for degradation, and to procedures for increasing the range of substrates that can be accepted. The ortho-substituted congeners are particularly recalcitrant. a. It has been suggested that the substrate specificity of the dioxygenase is determined by the α-subunit of the dioxygenase. In contrast, the substrate specificities of the four chimeras constructed from the respective α- and β-subunits of the terminal dioxygenase ISPBPH of Pseudomonas sp. strain LB400 and Comamonas (Pseudomonas) testosteroni strain B-356 were dependent on the presence of both proteins (Hurtubise et al. 1998). In addition, the catalytic activity of hybrid dioxygenases comprising α- and β-subunits from distant biphenyl dioxygenases is not determined specifically by one or other of the subunits (Chebrou et al. 1999). There is therefore a complex dependency on the presence of both subunits. b. Shuffling the genes in a fragment of the bphA gene in Burkholderia sp. strain LB400, Comamonas testosteroni strain B-356, and Rhodococcus globerulus strain P6 resulted in variants that had high activity toward the generally persistent 2,6-dichloro- and 2,4,4′trichlorobiphenyls (Barriault et al. 2002). c. Site-directed mutants of amino acids that coordinate the catalytic iron center of Pseudomonas pseudoalcaligenes strain KF707 were produced and expressed in Escherichia coli. One mutant was able to degrade (25-25) 2,5,2′,5′-tetrachlorobiphenyl by 3,4-dioxygenation and displayed 2,3- and 3,4-dioxygenase activities for 2,5,2′- and 2,5,4′-trichlorobiphenyls (Suenaga et al. 2002). d. Efforts have been made to increase the range of substrates accepted by the dioxygenase by mutagenesis of multiple sites in region III (Barriault and Sylvestre 2004). The changes were in the positions at which dioxygenation took place: • 2-2 (2,2′)-Dichloro from the 2:3 position with loss of chloride, to the 3,4-dihydrodiol that is a terminal metabolite • 23-23 (2,2′,3,3′)-Tetrachloro to the 4,5-dihydrodiol without dechlorination • 2,2′,5,5′-Tetrachloro to the 3,4-dihydrodiol without dechlorination e. An attractive suggestion has been made to improve the spectrum of PCBs that are degraded by Pseudomonas sp. strain KKS102 by altering the promoter of the bph operon

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(Ohtsubo et al. 2003). By this means, strains were obtained that showed enhanced degradation of tri-, tetra-, and pentachlorobiphenyl. The next step in degradation is generally dehydrogenation that results in the production of the dihydroxybiphenyl, which is the substrate for ring fission. Alternatively, in the degradation of 2,4′- and 4,4′-dichlorobiphenyl by Pseudomonas testosteroni strain B-356 involving 3,4-dioxygenation, this is accomplished by loss of chloride (Ahmad et al. 1991), as in Burkholderia (Pseudomonas) sp. strain LB400 with congeners carrying ortho-chlorine substituents (Haddock et al. 1995; Seeger et al. 1999). Ring Fission by 2,3-Dihydroxybiphenyl Dioxygenase Degradation of the initially formed 2,3-dihydroxybiphenyl is carried out by extradiol fission, and 2,3-dihydroxybiphenyl 1,2-dioxygenase has been purified and characterized from several organisms including Pseudomonas pseudoalcaligenes strain KF707 (Furukawa and Arimura 1987), Pseudomonas paucimobilis strain Q1 (Taira et al. 1988), Burkholderia (Pseudomonas) sp. strain LB400 (Haddock et al. 1995), and the Gram-positive Rhodococcus sp. strain RHA1 (Hauschild et al. 1996). In Rhodococcus globerulus strain P6 (Acinetobacter sp. strain P6), there are several nonhomologous 2,3-dihydroxybiphenyl-1,2-dioxygenases with a narrow substrate specificity (Asturias and Timmis 1993), and one of them encoded by bphC2 was appreciably different from other extradiol dioxygenases (Asturias et al. 1994). The enzyme from Bacillus sp. strain JF8 is unusual in being Mn(II)-dependent and differs in structure from those of Burkholderia sp. strain LB400 and Pseudomonas paucimobilis strain KF707 (Hatta et al. 2003). Detailed investigations have revealed a number of important conclusions that affect the degradability of PCB congeners: 1. In Burkholderia sp. strain LB400, which is one of the most versatile, dioxygenation at C-2 and C-3 is preferred, and hydroxylation at atoms bearing chlorine substituents is restricted to the ortho positions with elimination of chloride in place of dehydrogenation (Seeger et al. 1999). The preferred site also depends, however, on the substitution pattern of the other ring, and 2′- and 2′,6′-dichlorobiphenyls seriously inhibit the activity of dihydroxybiphenyl dioxygenase (Dai et al. 2002). It had been suggested that substrate inhibition of the dioxygenase was a result of enzyme inactivation, and a general mechanism for inhibition of the dioxygenase by catechols involving formation of the complex between the substrate and the Fe(II) enzyme has been suggested (Vaillancourt et al. 2002). This has been explored further (Fortin et al. 2005) using extradiol dioxygenases from organisms with established ability to degrade PCBs—Burkholderia sp. strain LB400, two isoenzymes from Rhodococcus globerulus strain P6, and Sphingomonas sp. strain RW1. Rates of dioxygenation were measured from the rate of formation of the fission product 2-hydroxy-6-keto-6phenylhexa-2,4-dienoate. Important conclusions emerged: (i) rates for the monochlorinated isomers were lower than those for the nonchlorinated 2,3-dihydrodihydroxybenzoate and (ii) among polychlorinated congeners, the rate for the recalcitrant 2′,6′dichloro-dihydroxybiphenyl was the lowest. 2. The next step in degradation involves hydrolysis of 2-hydroxy-6-keto-6-phenylhexa-2, 4-dienoate to benzoate and 2-hydroxy-penta-2,4-dienoate encoded by the hydrolase BphD. A study of the hydrolase in Burkholderia cepacia LB 400 has revealed significant details (Seah et al. 2000) (Figure 9.6a): a. There are substantial differences in the rates of hydrolysis by the hydrolase. b. The 4-chloro compound is hydrolyzed nonenzymatically to the 4-hydroxy compound, which may undergo cleavage to acetophenone and 3-ketoadipate (Figure 9.6b). c. The 3-chloro compound is stable to enzymatic hydrolysis and strongly inhibits the hydrolase.

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Environmental Degradation and Transformation of Organic Chemicals OH

OH

(a)

O O H CO 2H

OH

CO 2H

O O H CO 2H

HO

Cl

Cl

OH 2 5 8 9

(b) Cln

Cln

6

Cl

Co 2H

OH

4 3 2 O O H CO 2H 1

O O H CO 2 H Cln

10 OH

O O H CO 2H Cln O CH3 HO 2C CH2 O O H CO 2H

O

COO−CO 2H

Cln

FIGURE 9.6 Aerobic degradation of (a) biphenyl by biphenyl-2,3-dioxygenase, (b) a polychlorinated 4′-chlorobiphenyl.

3. The hydrolase is competitively inhibited by 3-chloro- and 4-chloro-2-hydroxy-6-keto6-phenylhexa-2,4-dienoate (Seah et al. 2000), and a cardinal role has been assigned to a glutathione S-transferase BphK that had not hitherto been assigned a specific function. The bphK gene does not occur in all bacteria that are recognized as being able to degrade PCBs, and it has been shown in Burkholderia (Pseudomonas) xenovorans strain LB400 that the transferase has the ability to dechlorinate 3-chloro-, 5-chloro-, and 3,9,11trichloro-2-hydroxy-6-keto-6-phenylhexa-2,4-dienoate in reactions involving addition of GS−, elimination of Cl−, followed by generation of GSSG (Fortin et al. 2006). These observations contribute to understanding the relative recalcitrance of PCB congeners, rationalize the formation of the metabolites, and focus attention on the formation of the 3- and 4-chlorohexadienoates that would be produced from 3- and 4-chlorinated biphenyls by 2-hydroxy-6-keto-6-phenylhexa-2,4-dienoate hydrolases. Collectively, these conclusions are supported by the results of a study of 33 congeners using Acinetobacter sp. strain P6 (Rhodococcus globerulus) (Asturias and Timmis 1993). The extradiol

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463

fission product greatly exceeded the chlorobenzoates for the 2,5,4′- and 2,4,4′-trichlorobiphenyls (Furukawa et al. 1979). In this strain, however, although the specificities of the hydrolases in Burkholderia cepacia LB 400 and Rhodococcus globerulus P6 were essentially similar, significant differences have been observed (Seah et al. 2001). In the former, inhibition by 3-chlorohexadienoate exceeded that of 4-chlorohexadienoate, whereas the opposite was found in the latter. Metabolites Chlorobenzoates and Catechols a. Total degradation of PCBs necessitates degradation of the chlorobenzoates produced by the foregoing reactions. It has been suggested that the inability of strains to degrade chlorobenzoates produced from some PCB congeners may be related to the restricted metabolism of chlorobenzoate fission products (Hernandez et al. 1995). The distribution of chlorobenzoates has indeed been suggested as biomarkers for the aerobic degradation of PCBs (Flanagan and May 1993). The inhibitory nature of chlorocatechols that are metabolites of chlorobenzoates has already been noted. b. Another metabolic limitation involves the formation of protoanemonin (4-methylenebut2-ene-4-olide) as an intermediate in the degradation of 4-chlorobenzoate formed by partial degradation of 4-chlorobiphenyl. This is formed by the intradiol fission of 4-chlorocatechol to 3-chloro-cis,cis-muconate followed by loss of CO2 and chloride (Blasco et al. 1995) (Figure 9.7). The synthesis of this metabolite adversely affects the survival of organisms that metabolize 4-chlorobiphenyl in soil microcosms, although its formation can be obviated by organisms using a modified pathway that produces the cis-dienelactone (Blasco et al. 1997). Acetophenones Quite different metabolites may also be formed. Although the ultimate products from the degradation of PCBs are generally chlorinated muconic acids, the unusual metabolite 2,4,5-trichloroacetophenone has been isolated (Bedard et al. 1987b) from the degradation of 2,4,5,2′,4′,5′-hexachlorobiphenyl (Figure 9.8) by Alcaligenes eutrophus H850, which has an unusually wide spectrum of degradative activity for PCB congeners. The mechanism for the formation of 2,4,5-trichloroacetophenone may be analogous to the formation of acetophenones by nonenzymatic hydrolysis of the 4′-chlorinated compound. The metabolism of chloroacetophenones takes place by monooxygenation to the corresponding chlorophenylesters. Some ortho-substituted chlorophenols are, however, inhibitory, and only low rates of oxidation have been encountered with di- and trichlorinated

O CO 2 H CO 2 H

O

OH OH

O

CO 2 H

CH 3 CO 2 H Cl

Cl

Cl

O CO 2 H O O CO 2 H

FIGURE 9.7

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CO 2 H

Formation of protoanemonin from 4-chlorobenzoate.

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Environmental Degradation and Transformation of Organic Chemicals Cl Cl

Cl Cl

Cl Cl

Cl COCH 3

Cl Cl

FIGURE 9.8

Biodegradation of 2,4,5,2′,4′,5′-hexachlorobiphenyl.

acetophenones, so that growth with them is not possible (Higson and Focht 1990). However, degradation of 4-chloroacetophenone has been demonstrated with a mixed culture of an Arthrobacter sp. and a Micrococcus sp. (Havel and Reineke 1993). The NIH Shift A rearrangement (NIH shift) occurred during the transformation of 2-chlorobiphenyl to 2-hydroxy3-chlorobiphenyl by a methanotroph, and is consistent with the formation of an intermediate arene oxide (Adriaens 1994). The occurrence of such intermediates also offers plausible mechanisms for the formation of nitro-containing metabolites that have been observed in the degradation of 4-chlorobiphenyl in the presence of nitrate (Sylvestre et al. 1982). Induction of PCB Metabolism

Apart from induction of degradative enzymes by growth with biphenyl, nonrelated naturally occurring substrates have been shown to induce the enzymes for PCB degradation. a. A range of related compounds was examined (Donnelly et al. 1994) for their capacity to support the growth of Alcaligenes eutrophus H850, Pseudomonas putida LB400, and Corynebacterium sp. MB1. For strains H850 and MB1, growth with biphenyl equaled that using a wide range of substrates including naringin, catechin, and myricitin (Figure 9.9). These results suggest that natural plant metabolites are able to mediate the growth of PCBdegrading organisms. In addition, the pattern of metabolism of PCB congeners was identical using biphenyl or the naturally occurring plant metabolites. b. Cells of Arthrobacter sp. strain B1B were grown in a mineral medium with fructose and carvone (50 mg/L). Effective degradation of a number of congeners in Arochlor 1242 was induced by carvone that could not, however, be used as a growth substrate, and was toxic at high concentrations (>500 mg/L). Other structurally related compounds including limonene, p-cymene, and isoprene were also effective (Gilbert and Crowley 1997). These results may be particularly relevant in the context of bioremediation, and lend support to the potential role of the rhizosphere and of plants (Chapters 12 and 14, Part 4). There are other changes that are involved in the metabolism of biphenyls and PCBs. A study of the degradation of biphenyl by Burkholderia xenovorans LB400 showed that elevated levels of the enzymes involved in the upper Bph pathway were induced by growth with biphenyl, whereas those for benzoate degradation were upregulated by growth on benzoate. Enzymes for the latter involved conversion to benzoyl-CoA and use of the pathway delineated for Azoarcus evansii (Denef et al. 2004). Accumulation of inorganic polyphosphate is induced in cells exposed to stress, and has been observed during the growth of Pseudomonas sp. strain B4 in a defined medium with biphenyl, and on changing the growth substrate from glucose to biphenyl (Chávez et al. 2004).

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465 OH

O

HO

OH

HO

O

O

HO

Esculetin

O Quercitin

OH OH

OH

OH

O

O

OH HO

OH

O

HO

Myricitin

O Naringenin

OH

OH

OH OH OH

HO

O

O Catechin

OH OH

FIGURE 9.9

Carvone

Limonene

Naturally occurring products used as substrates or inducers.

FUNGAL DEHALOGENATION Dehalogenation of commercial PCB mixtures has been observed using Phanerochaete chrysosporium in different media including low nitrogen, high nitrogen, and complex malt extract (Yadav et al. 1995). Dechlorination of congeners with ortho, meta, and para substituents occurred. Maximal dechlorination ranged from 61% for Arochlor 1242 to 18% for Arochlor 1260, and although it was greatest in malt extract medium, it also occurred in low- and high-nitrogen media.

REDUCTIVE DEHALOGENATION The anaerobic dechlorination of PCBs has been extensively studied both in laboratory microcosms and in field samples from heavily contaminated sites in the United States. A complex pattern of dechlorination has been described that depends on the site and environmental factors such as concentration of organic carbon and sulfate (Bedard and Quensen 1995). In a study where dechlorination was stimulated by addition of brominated biphenyls (Bedard et al. 1998; Wu et al. 1999), three main patterns were found—N that removed flanked meta chlorines, P that removed para chlorines, and LP that removed unflanked para chlorines. The example of the 2,2′,3,4,4′,5,5′-heptachloro (2345-245) congener is given in Figure 9.10. For the heptachloro congener 2,2′,3,3′,4,4′,5 (2345234), a combination of processes N, P, and LP resulted in the production of the 2,2′-dichloro- and 2,2′,5-trichloro congeners (Bedard et al. 2005). In contrast, ortho chlorines were more recalcitrant. An important study using sediment contaminated with Arochlor 1260 was initiated by priming with 2,6-dibromobiphenyl to select for N dechlorination. Successive dilution produced sedimentfree cultures that grew with acetate, butyrate or pyruvate, and H2 and resulted in a range of bacteria. These were able to dechlorinate hexa- through nonachlorinated congeners to tri- through

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Environmental Degradation and Transformation of Organic Chemicals 2345-245-CB

Process N

Process P

Cl

Cl

Cl

Cl

Cl

Cl Cl

Cl

Cl 235-25-CB

24-24-CB

Cl

Cl 2-2-CB

FIGURE 9.10 Anaerobic dechlorination of 2,2′,3,4,4′,5,5′-heptachlorobiphenyl. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.)

pentachlorinated congeners, and dechlorination of heptachloro congeners produced the 2,2′,4′,5 (25-24), 2,2′,4,6′ (24-26) and 2,2′,5,6′ (25-26) tetrachloro congeners with retention of the ortho substituents (Bedard et al. 2006). It is important to appreciate important factors that affect the relative effectiveness of dechlorination: (a) temperature (Wu et al. 1997b) that is discussed in greater detail below and (b) the nature of the PCB congener that primes the dehalogenation (Wu et al. 1997a). In contrast to the results discussed above, an enrichment culture supplemented with acetate was able to dechlorinate 2,3,5,6-tetrachlorobiphenyl to 3,5-dichlorobiphenyl with the removal of ortho chlorines (Holoman et al. 1998), and success is being achieved in defining the organisms responsible. For example, a nonmethanogenic mixed culture containing strain o-17 was able to bring about dechlorination of ortho-halogenated congers 2,3,5,6-tetrachlorobiphenyl to 3,5-dichlorobiphenyl (May et al. 2006). Substantial attention has been given to organisms within the Dehalococcoides group and the green nonsulfur phylum Chloroflexi (Fagervold et al. 2005). Examples include the following: (a) laboratory studies using enriched cultures specifically dechlorinated PCBs with doubly flanked chlorines (Wu et al. 2002), (b) cultures containing Dehalococcoides ethenogenes carried out sequential dechlorination of 2,3,4,5,6-pentachlorobiphenyl to 2,4,6-trichlorobiphenyl (Fennell et al. 2004), and (c) it has been shown that individual species of Chloroflexi have different specificities (Fagervold et al. 2005). The Role of Temperature This is an important parameter particularly for naturally occurring mixed cultures of organisms in the natural environment: temperature may result in important changes in the composition of the microbial flora as well as on the rates for different processes. An illustrative example of its importance includes the following. An anaerobic sediment sample was incubated with

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Substituted Carbocyclic Aromatic Compounds

467

2,3,4, 6-tetrachlorobiphenyl at various temperatures between 4 and 66°C (Wu et al. 1997a,b). The main products were 2,4,6- and 2,3,6-trichlorobiphenyl and 2,6-dichlorobiphenyl; the first was produced maximally and discontinuously at 12 and 34°C, the second maximally at 18°C, and the third was dominant from 25 to 30°C. Dechlorination was not observed above 37°C. Collectively, there is therefore extensive evidence for the anaerobic dechlorination of PCBs but the extent and specificity of this depends on both the organism and the degree and pattern of substituents in the congener. Polybrominated Biphenyls and Diphenylmethanes

Mixed cultures of organisms that were isolated from sediments contaminated with PCBs and polybrominated biphenyls (PBBs) were shown to debrominate PBBs under anaerobic conditions (Morris et al. 1992), and the dominant congener—2,2′,4,4′,5,5′(245-245)-hexabromobiphenyl—could be successively debrominated to 2,2′-dibromobiphenyl. However, in sediments from the most heavily contaminated site containing contaminants in addition to PBBs, very little debromination occurred and the recalcitrance was attributed to the toxicity of the other contaminants (Morris et al. 1993). Ingenious experiments to which reference has already been made have used addition of specific PCB congeners that are more readily dechlorinated to “prime” dechlorination at specific positions (Bedard and Quensen 1995). They have been extended to the use of dibrominated biphenyls in the presence of malate to stimulate dechlorination of the hexachloro- to nanochlorobiphenyls (Bedard and Quensen 1995). Results from these experiments provide valuable evidence of important differences between anaerobic dechlorination and anaerobic debromination, and the greater facility of the latter. The use of brominated biphenyls to induce dechlorination of highly chlorinated biphenyls has been examined in detail. Di- and tribromobiphenyls were the most effective in dechlorinating the PCBs including the heptachloro, hexachloro, and pentachloro congeners, and were themselves reduced to biphenyl (Bedard et al. 1998). In addition, 2,6-dibromobiphenyl stimulated the growth of anaerobes that effectively dechlorinated hexa-, hepta-, octa-, and nanochlorobiphenyls over the temperature range 8–30°C (Wu et al. 1999). In anaerobic sediment microcosms, a range of tribrominated biphenyls was successively debrominated to dibromo- and monobromo compounds before complete debromination to biphenyl. Of particular interest, both the 2,2′,4′,5 (25-24) and 2,2′,5,5′ (25-25) tetrabromo congeners were debrominated to the 2,2′-dibromo congener that was slowly and completely debrominated to biphenyl within 54 weeks (Bedard and van Dort 1998). The pathways for debromination of 2,2′,4,5′- and 2,2′,5,5′-tetrabromobiphenyls are shown in Figure 9.11. In comparable microcosm experiments, a number of important features were observed: a. All the tribrominated congeners were debrominated to products including biphenyl in acclimation times between <1 and 7 weeks. b. For the corresponding chlorinated congeners with comparable acclimation times, the loss of only single chlorine was noted and biphenyl was not observed. c. The mono- and dichlorinated congeners required long periods of acclimation before the onset of dechlorination. Other compounds have also been examined as “primers” for the dechlorination of hexachloro to nonachloro PCB congeners (DeWeerd and Bedard 1999): a number of substituted brominated monocyclic aromatic compounds were examined, and 4-bromobenzoate was effective—though less so than 2,6-dibromobiphenyl. In contrast, the chlorobenzoates that are metabolites of aerobic degradation were ineffective. The positive effect of brominated biphenyls in “priming” the anaerobic dechlorination of CBs has also been encountered in the dechlorination of octachlorodibenzo[1,4] dioxin to the 2,3,7,8 congener induced by 2-bromodibenzo[1,4]dioxin in the presence of H 2 (Albrecht et al. 1999).

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Environmental Degradation and Transformation of Organic Chemicals Br

Br

Br

Br Br Br

Br 2 4 -2 5 -B B Br

Br 2 4 -2 -B B Br

Br

Br Br 2-2-BB

Br

Br 2 5 -2 5 -B B

Br 2 5 -2 -B B Br

2-BB

Biphenyl

FIGURE 9.11 Anaerobic debromination of 2,2′,4′,5(24-25) and 2,2′,5,5′(25-25)-tetrabromobiphenyl. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.)

PBBs are therefore not only debrominated microbially under anaerobic conditions, but are also able to induce effective dechlorination of their chlorinated analogs. Debromination may, however, be limited in the presence of other contaminants. Polyhalogenated Dioxins and Diphenyl Ethers Anaerobic dehalogenation has also been observed in aromatic ethers including polychlorinated dibenzo[1,4]dioxins (Fennell et al. 2004; Yoshida et al. 2005). The anaerobic debromination of decabromodiphenyl ether is considered here since it is analogous to the dehalogenation of polyhalogenated biphenyls. a. Debromination took place with formation of octa- and nanobrominated congeners, predominantly by loss of bromine at the meta and para positions, and not appreciably at the ortho position. In addition, the rate was increased by the addition of primers (Gerecke et al. 2005). b. Another approach used two anaerobes that were able to dechlorinate chlorinated ethenes, and the results showed significant differences between them. Cells were grown with trichloroethene, and Sulfurospirillum multivorans debrominated decabromodiphenyl ether to octa and hepta congeners. In contrast, a strain of Dehalococcoides sp. (ANAS) that was unable to debrominate decabromodiphenyl ether was able to carry out a range of debrominations of the octa congeners to the hexa (154) and, subsequently, to the penta (99) and tetra (49) + tetra (47) congeners (He et al. 2006).

HALOGENATED BENZOATES Chlorobenzoates may be formed during the initial steps in the aerobic degradation of PCBs, and their further metabolism illustrates a number of pathways. There are several reactions that carry out dehalogenation including dioxygenation, hydrolysis, and reduction.

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Substituted Carbocyclic Aromatic Compounds

469 OH CO 2 H COCl

CO 2 H

2:3

OH OH

Cl

Cl 1:2

CO 2 H CO 2 H Cl

OH CO 2 H COH

OH

CHO

2:3

Cl−CH 2CHO

OH Cl

Cl

Cl

1:2 Cl

CO 2 H CO 2 H

FIGURE 9.12 Ring-cleavage pathways for the biodegradation of chlorocatechols formed from 3- and 4-chlorobenzoates.

Dioxygenation The aerobic degradation of benzoate is initiated by dioxygenation with the formation of catechol and concomitant decarboxylation. An analogous pathway is used for halogenated benzoates with ortho substituents, which involves dioxygenation with decarboxylation and loss of halide. For example, a broad-spectrum two-component 1,2-dioxygenase is used by Pseudomonas cepacia 2CBS for 2-halobenzoates, and a three-component 1,2-dioxygenase for 2-chloro- and 2,4-dichlorobenzoate by Pseudomonas aeruginosa strain 142 (Romanov and Hausinger 1994). The alternative 2,3-dioxygenation to give 2,3-dihydroxybenzoate with loss of chloride has been observed in a Pseudomonas sp. (Fetzner et al. 1989). For 3- and 4-halogenated benzoates, dioxygenation and dehydrogenation with concomitant decarboxylation without loss of halogen is generally observed. For example, 3-chlorobenzoate may produce 3-chlorocatechol by 1,2-dioxygenation or 4-chlorocatechol by 1,6-dioxygenation, and 4-chlorobenzoate would produce 4-chlorocatechol. The pathway for the further metabolism of the chlorocatechols may be critical due to the formation of toxic metabolites by extradiol (2:3) fission— an acyl chloride from 3-chlorocatechol or chloroacetaldehyde from 4-chlorocatechol (Figure 9.12). The “modified ortho” (1:2) fission may exceptionally take place, and has already been noted. 1. A two-component 2-halobenzoate 1,2-dioxygenase has been purified from Pseudomonas cepacia strain 2CBS that is able to metabolize 2-fluorobenzoate, 2-chlorobenzoate, 2-bromobenzoate, and 2-iodobenzoate to catechol by concomitant decarboxylation and loss of halide (Fetzner et al. 1992). The inducible 2-halobenzoate 1,2-dioxygenase consisted of

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Environmental Degradation and Transformation of Organic Chemicals O

CO 2 H

O.C Br

CO 2 H CO 2 H

OH

OH OH Br

OH

CO 2 H CO 2 H

O O HO 2 C

O

FIGURE 9.13 Degradation of 2-bromobenzoate by Pseudomonas aeruginosa strain 142. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.)

two components: an oxygenase A and a reductase B. The oxygenase A is an iron–sulfur protein and consists of nonidentical subunits, α (Mr 52,000) and β (Mr 20,000) in an α3β3 structure. The reductase B is an iron–sulfur flavoprotein (Mr 37,500) containing FAD. It is broadly similar to benzoate dioxygenase though different from the FMN-containing dioxygenases. It has been shown that the gene cluster cdbABC encoding the enzyme is localized on a plasmid (Haak et al. 1995). Significant homology existed with the amino acid sequence of benABC, which encodes benzoate dioxygenase in Acinetobacter calcoaceticus and to xylXYZ, which encodes toluate dioxygenase in Pseudomonas putida mt-2. 2. The 2-halobenzoate dioxygenase of Burkholderia sp. strain TH2 was active toward a number of 2-halobenzoates, and the predicted amino acid sequences of all the gene products cbdABC were highly similar to those of Pseudomonas sp. strain 2CBS. It was shown that the effectors of the transcriptional regulatory gene Cbds were 2-chloro-, 2-bromo-, 2-iodo-, and 2-methylbenzoate (Suzuki et al. 2001). 3. Pseudomonas aeruginosa strain 142 was isolated from a PCB-degrading consortium and is able to degrade a range of 2-halogenated benzoates (Figure 9.13). The enzyme consists of three components, one of which is a ferredoxin containing a single Rieske-type [2Fe–2S] cluster (Romanov and Hausinger 1994), and the amino acid sequence is similar to those of other three-component dioxygenases containing ferredoxin such as benzene dioxygenase, toluene-2,3-dioxygenase, biphenyl dioxygenase, and naphthalene dioxygenases. This pathway is used for the degradation of both 2,3- and 2,5-dichlorobenzoate (Hickey and Focht 1990), and for a number of other 2-substituted benzoates (Romanov and Hausinger 1994). 4. Burkholderia sp. strain NK8 can be grown with benzoate, and both 3- and 4-chlorobenzoate. The genes encoding enzymes for the dioxygenation of benzoates (cbeABCD) and catechols (catA, catBC) have been cloned and analyzed (Francisco et al. 2001). Both 3-chlorocatechol (1:6 dioxygenation) and catechol (1:2 dioxygenation) were formed from 2-chlorobenzoate, and 4-chlorocatechol (1:2 dioxygenation) from both 3- and 4-chlorobenzoate. Degradation of both the 3- and 4-chlorocatechol was accomplished by intradiol fission. The degradation of muconates derived from 3- and 4-catechol by intradiol fission mediated by the 1,2-dioxygenase is well established (Figure 9.14a,b,c), and involves the activity of a sequence of enzymes: chloromuconate cycloisomerase, dienelactone hydrolase, and maleylreductase. One of the key steps in degradation occurs after ring fission, and involves maleylreductase. For maleylacetates with a halogen atom at the 2-position, reductive loss of halide takes place using 2 mol of NADH per mol of substrate, whereas only 1 mol of NADH is required for those lacking a halogen substituent (Figure 9.15) (Kaschabek and Reineke 1995). For 3,5-dichlorocatechol that would be produced from

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Substituted Carbocyclic Aromatic Compounds

471

OH OH

CO 2H CO 2H

(a)

HO 2 C

HO 2 C

O

O

HO 2C

CO 2H CO 2 H

OH

HO 2 C

O

O

O

HO 2 C

O

CO 2 H

O

O

O

Cl

OH

Cl

(b)

Cl

Cl HO 2 C

HO 2 C

O

HO 2 C

O

CO 2H

O CO 2 H

CO 2 H

Cl

HO 2C

CO 2 H CO 2 H

HO 2 C

O

O

O

O

OH OH

Cl

Cl

Cl

(c)

HO 2 C

Cl

HO 2 C

O

O

CO 2H

CO 2 H

Cl

FIGURE 9.14

Degradation of (a) catechol, (b) 3-chlorocatechol, and (c) 4-chlorocatechol.

O

O

HO 2 C

O

Hal

O

HO 2 C

HO 2 C CH 2 CO 2 H

Hal Di enelactone

CH 2 CO 2 H

CO 2 H CO 2 H

H Maleylacetate

O β-Ketoadipate

FIGURE 9.15 Degradation of halogenated dienelactones maleylacetate and reduction to β-ketoadipate. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.)

3,5-dichlorobenzoate, or 2,4-dichlorophenol, from 2,4-dichlorophenoxyacetate, there is, however, some ambiguity over the details of the formation of the final products (succinate and acetate). 5. Alcaligenes sp. strain BR6024 degraded 3-chlorobenzoate by dioxygenation to 3-chlorobenzoate-cis-4,5-dihydrodiol, dehydrogenation to 5-chloro-3,4-dihydroxybenzoate, and formation of pyr-2-one-2,4-dicarboxylate and 3,4-dihydroxybenzoate (Nakatsu et al. 1997). The cbaC gene that encodes 3-chlorobenzoate-cis-4,5-dihydrodiol dehydrogenase is not,

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however, required for the metabolism of 3,4-dichlorobenzoate when spontaneous loss of halide occurs (Nakatsu et al. 1997). 6. For 3-chlorobenzoate, an alternative pathway in Alcaligenes sp. strain BR60 may involve 3,4- or 4,5-dioxygenation. Ring fission of the catechols resulted in the production of pyruvate and oxalacetate (Nakatsu and Wyndham 1993). These considerations are clearly relevant to the degradation of PCBs in which chlorobenzoates are produced during the initial dioxygenation of the upper pathway. The products from metabolism of these, such as 3-chlorocatechol, may inhibit the function of the ring-fission dioxygenase (Sondossi et al. 1992), while the synthesis of protoanemonin from 4-chlorobenzoate may inhibit the survival of organisms that degrade 4-chlorobiophenyl (Blasco et al. 1997). Hydrolytic Reactions 4-Halobenzoates The metabolism of 4-halobenzoates is different from that of 2-halobenzoates since simultaneous loss of halide and the carboxyl group cannot occur. Instead, hydrolytic loss of halogen is involved as an early step in the degradation of 4-halogenated benzoate by a number of strains. It has been shown that the carboxyl group of strains with hydrolytic dehalogenating activity is generally activated by a coenzyme A ligase to form benzoyl-coenzyme A esters. The substrate specificity of the 4-halobenzoate-CoA ligase in Pseudomonas sp. strain CBS3 (Löffler et al. 1992) and the kinetic parameters for dehalogenation of 4-halobenzoyl CoA by Acinetobacter sp. 4-CB1 (Crooks and Copley 1993) have been provided. 1. Alcaligenes denitrificans strain NTB-1 is able to use 4-chloro-, 4-bromo-, and 4-iodobenzoates as sole sources of carbon and energy. The pathway involves hydrolytic dehalogenation to 4-hydroxybenzoate followed by hydroxylation to 3,4-dihydroxybenzoate (van den Tweel et al. 1987). 2. A comparable pathway was used by Arthrobacter sp. strain TM-1, which was able to grow with 4-chlorobenzoate and 4-bromobenzoate (Marks et al. 1984). 3. Hydrodehalogenation of 4-chlorobenzoate has been shown in cell extracts of Pseudomonas sp. strain CBS3 (Löffler et al. 1991), and the resulting 4-hydroxybenzoate is then readily degraded by hydroxylation to 3,4-dihydroxybenzoate followed by ring fission. The dehalogenase consists of three components, one with a molecular weight of 3000 that is unstable, and two stable components with molecular weights of ≈ 86,000 and 92,000 (Elsner et al. 1991; Chang et al. 1992). 4. A strain of Acinetobacter sp. 4-CB1 is able to dehalogenate 4-chlorobenzoate with the formation of 4-hydroxybenzoate after the initial formation of the 4-chlorobenzoyl CoA ester (Copley and Crooks 1992). An additional variant that involves both hydrolytic and reductive reactions has been found in the degradation of 2,4-dichlorobenzoate by Alcaligenes denitrificans strain NTB-1 (van den Tweel et al. 1987) (Figure 9.16), and via the CoA ester by Corynebacterium sepedonicum (Romanov and Hausinger 1996).

MECHANISMS FOR THE RING FISSION OF SUBSTITUTED CATECHOLS It is convenient to provide a short introduction to the issues encountered in the degradation of chlorinated substrates that produce chlorocatechols as intermediates. An outline of the ring-fission mechanisms has been given in Chapter 3, Part 1, and Chapter 8, Part 3. Three pathways for fission

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Substituted Carbocyclic Aromatic Compounds CO 2 H

473

CO 2 H

CO 2 H

Cl

OH

CO 2 H

Cl

OH Cl

FIGURE 9.16

OH

Reductive step in the biodegradation of 2,4-dichlorobenzoate.

of 3-substituted catechols by dioxygenation are possible—1,2-intradiol (ortho), 2,3-extradiol (meta), and 1,6- (meta/distal) (Figure 9.2). The rather complex situation may be summarized as follows: 1. Extradiol fission of substituted catechols is generally favored except for 3-chlorocatechol when the enzyme is inhibited by the substrate (Klecka and Gibson 1981; Bartels et al. 1984). Intradiol fission is then favorable using what has been termed a “modified ortho pathway.” After growth of Pseudomonas chlororaphis strain RW71 with 1,2,3,4-tetrachlorobenzene, the chlorocatechol 1,2-dioxygenase is able to catalyze the oxidation of a wide range of chlorocatechols (Potrawfke et al. 1998). Intradiol fission has been observed even for 4,5-dichlorocatechol, although this is not considered to be normally accepted by chlorocatechol dioxygenases (Potrawfke et al. 2001). An important implication is that the degradation of 3-methylcatechol, for which extradiol fission is preferred, may be incompatible with the simultaneous degradation of 3-alkyl and 3-halogenated catechols. 2. Pseudomonas putida strain GJ31 degrades chlorobenzene via 3-chlorocatechol and extradiol fission (Kaschabek et al. 1998). This is accomplished by a chlorocatechol 2,3dioxygenase that hydrolyzes the initially formed cis,cis-hydroxymuconacyl chloride to 2-hydroxymuconate. Thereby, the irreversible reaction of the acid chloride with nucleophiles or the formation of pyr-2-one-6-carboxylate as a terminal metabolite is avoided (Mars et al. 1997). The 3-chlorocatechol 2,3-dioxygenase has been characterized, and is clearly different from other catechol extradiol dioxygenases (Kaschabek et al. 1998; Mars et al. 1999). An analogous spontaneous hydrolysis has been observed in the product from 1:2 dioxygenation of 2-chlorohydroquinone, which is a metabolite of γ-hexachlorocyclohexane (Endo et al. 2005). 3. The alternative extradiol fission between C-1 and C-6 (distal fission) of catechol has been observed in Azotobacter vinelandii strain 206 (Sala-Trepat and Evans 1971). Although this is rather inefficient due to the low turnover capacity for the enzyme, it has been observed for 3-chlorocatechol in Sphingomonas xenophaga (Riegert et al. 1998), and its catalytic properties have been improved by random mutagenesis of bphC1 that encodes 2,3-dihydroxybiphenyl dioxygenase (Riegert et al. 2001). Which of the pathways is followed depends therefore on both the organism and the substrate that is being metabolized. In summary, a range of pathways are available for the degradation of halogenated benzoates: • Dioxygenation with concomitant loss of carboxyl and halogen to produce catechols (2-, 2,3-, and 2,5-halogenated benzoates) • Dioxygenation with decarboxylation but without loss of halogen to produce chlorocatechols that may be degraded by several ring-fission pathways (3- and 4-chlorobenzoates) • Dioxygenation with loss of only halogen to produce a dihydroxybenzoate (2-chlorobenzoate)

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Environmental Degradation and Transformation of Organic Chemicals R

R

R H OH OH H

OH

OH

R O CO 2H OH

FIGURE 9.17

CO−CO 2H

CH 3−CHO + CH 3−CO−CO2H

Biodegradation of alkylbenzenes.

• Dehalogenation of the CoA ester to hydroxybenzoate (4-halobenzoates) • Both hydrolytic and reductive elimination of halogen (2,4-dichlorobenzoate) There is an additional problem that has important implications for the bioremediation of contaminated sites when two substrates such as a chlorinated and an alkylated aromatic compound are present. The extradiol fission pathway is generally preferred for the degradation of alkylbenzenes (Figure 9.17), although this may be incompatible with the degradation of chlorinated aromatic compounds since the 3-chlorocatechol produced inhibits the activity of the catechol-2,3-oxygenase (Klecka and Gibson 1981; Bartels et al. 1984). This has been overcome in some strains: 1. Mutant strains have successfully reconciled this incompatibility (Taeger et al. 1988; Pettigrew et al. 1991). 2. 2,3-Dihydroxybiphenyl 1,2-dioxygenase from the naphthalene sulfonate–degrading Sphingomonas sp. strain BN6 metabolized 3-chlorocatechol by extradiol fission between the 1- and 6-positions (distal fission) (Riegert et al. 1998) 3. Chlorocatechol 2,3-dioxygenase from Pseudomonas putida GJ31 metabolized 3-chlorocatechol with concomitant elimination of chloride to form 2-hydroxymuconate (Kaschabek et al. 1998), while the catechol 2,3-dioxygenase from this strain encoded by cbzE is plasmid-borne and is capable of metabolizing both 3-chlorocatechol and 3-methylcatechol (Mars et al. 1999). It belongs to the 2.C subfamily of type 1 extradiol dioxygenases. 4. Attempts have been made to overcome this limitation by using random mutagenesis of the genes of the 2,3-dihydroxybiphenyl dioxygenase from Sphingomonas xenophaga (Riegert et al. 2001). This resulted in higher rates of reaction with 3-chlorocatechol and mutants that were able to degrade 3-methylcatechol and 2,3-dihydroxybiphenyl by distal fission. Reductive Loss of Halogen This is an important though less common reaction carried out by anaerobic bacteria, and may involve only partial dehalogenation; for example, 2,4,6-chlorobenzoate is successively dechlorinated by enrichment cultures to 2,4-dichlorobenzoate and 4-chlorobenzoate that is the terminal product (Gerritse et al. 1992). Although a number of strains were isolated by the enrichment of halogenated benzoates under denitrifying conditions, no growth on either 2- or 4-bromobenzoate was observed under either aerobic or denitrifying conditions. Some strains were able to grow with 3-bromobenzoate under denitrifying conditions but not aerobically. On the basis of 16S rRNA

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475

sequences, these organisms belonged to Thauera aromatica, Pseudomonas stutzeri, or Ochrobactrum anthropi (Song et al. 2000). Other illustrations of reductive pathways employed by aerobic organisms include the following: a. Alcaligenes denitrificans strain NTB-1 (now designated as a coryneform) that was isolated with 4-chlorobenzoate is able to grow with 4-bromo- and 4-iodobenzoate and, in addition, on 2,4-dichlorobenzoate. The first step in the pathway for 2,4-dichlorobenzoate is the reductive loss of the ortho chlorine substituent, whereas the second step is carried out by a halohydrolase that forms 4-hydroxybenzoate, followed by ring fission (Figure 9.16) (van den Tweel 1987). b. It has been shown that the reductive dechlorination of this strain and of Corynebacterium sepedonicum involves initial synthesis of a coenzyme A thioester and involves NADPH as the reductant. In addition, hydrolytic 4-chlorobenzoyl-CoA dehalogenase, 4-hydroxybenzoate 3-monooxygenase, and 3,4-dihydroxybenzoate 3,4-dioxygenase activities were found and enabled the construction of the degradative pathway (Romanov and Hausinger 1996). Reductive dehalogenation by anaerobic bacteria has been most extensively explored for halogenated phenols that are discussed in Part 2 of this chapter. Desulfomonile tiedjei was the first organism isolated in pure culture. It is a sulfate-reducing bacterium that is capable of reducing 3-chlorobenzoate to benzoate (DeWeerd et al. 1990) when ATP is synthesized by coupling proton translocation to dechlorination (Dolfing 1990; Mohn and Tiedje 1991). Cells induced by growth with 3-chlorobenzoate were able to partially dechlorinate polychlorinated phenols, specifically at the 3-position, whereas the monochlorophenols were apparently resistant to dechlorination (Mohn and Kennedy 1992). The membrane-bound reductive dehalogenase from Desulfomonile tiedjei has been solubilized and purified (Ni et al. 1995). It is distinct from the tetrachlorohydroquinone enzyme from a strain of Flavobacterium sp. (Xun et al. 1992c), and it plausibly plays a role in the energy transduction of Desulfomonile tiedjei. A membrane-bound cytochrome c is coinduced with the activity for reductive dechlorination, and has been purified and shown to be a high-spin diheme cytochrome distinct from previously characterized c-type cytochromes (Louie et al. 1997). It has been suggested that a chemiosmotic process may be used to rationalize the coupling of energy production with concomitant dechlorination (Louie and Mohn 1999). The related Desulfomonile limimaris is able to dechlorinate several meta-chlorinated benzoates including 2,3,5-trichlorobenzoate (Sun et al. 2001).

HALOGENATED PHENYLACETATES The degradation of 4-halogenated phenylacetates is carried out by 3:4 dioxygenation with loss of the halogen. Pseudomonas sp. strain CBS is able to grow at the expense of 4-chlorophenylacetate, and loss of chloride is mediated by a dioxygenase, which produces 3,4-dihydroxyphenylacetate that is degraded by an extradiol catechol dioxygenase (Markus et al. 1986). The dioxygenase consists of two components A and B. Component A has been purified and has a molecular weight of 140,000 consisting of three equal subunits, and contains iron and acid-labile sulfur. The substrate specificity is given in Table 9.1. Component B is a monomeric reductase with a molecular weight of 35,000 and contains per mol of enzyme, 1 mol of FMN, 2.1 mol of Fe, and 1.7 mol of labile sulfur. After reduction with NADH, the ESR spectrum showed signals that were attributed to a [2Fe–2S] structure and a flavosemiquinone radical (Schweizer et al. 1987). The molecular and kinetic properties of the enzyme are broadly similar to the Class IB reductases of benzoate 1,2-dioxygenase and 4-methoxybenzoate monooxygenase-O-demethylase.

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TABLE 9.1 Substrate Specificity of 4-Chlorophenylacetate 3,4-Dioxygenase in Component A of Pseudomonas sp. Strain CBS (Markus et al. 1986) Relative Activity (%)

Substrate 4-Fluorophenylacetate 4-Chlorophenylacetate

30 100

4-Bromophenylacetate 3-Chlorophenylacetate

CO 2H

CO 2 H I

I

R

R

Relative Activity (%)

Substrate

102 10

CO 2H

I

I

I

R

R

R

R

CO 2H I

I

R

R I

R=

NH−COCH 3

FIGURE 9.18 Transformation of 3,5-diacetamido-2,4,6-triiodobenzoate by Trametes versicolor. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.)

Fungal Reactions Substituted 2,4,6-triiodinated benzoates are incorporated into x-ray contrast agents, and their transformation has been examined in the white-rot fungus Trametes versicolor (Rode and Müller 1998). Since these compounds are putatively unable to pass the cell walls of the fungus, it is important that although lignin peroxidase activity was not observed, nonspecific extracellular manganese-dependent peroxidase and laccase activities were found. There was no introduction of oxygen into the ring and the main reactions were successive deiodination to the monoiodinated compound (Figure 9.18).

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Suzuki K, N Ogawa, K Miyashita (2001) Expression of 2-halobenzoate dioxygenase genes (cbdSABC) involved in the degradation of benzoate and 2-halobenzoate in Burkholderia sp. TH2. Gene 262: 137–145. Sylvestre M, R Massé, F Messier, J Fauteux, J-G Bisaillon, R Beaudet (1982) Bacterial nitration of 4-chlorobiphenyl. Appl Environ Microbiol 44: 871–877. Taeger K, H-J Knackmuss, E Schmidt (1988) Biodegradability of mixtures of chloro- and methylsubstituted aromatics: simultaneous degradation of 3-chlorobenzoate and 3-methylbenzoate. Appl Microbiol Biotechnol 28: 603–608. Taira K, J Hirose, S Hayashida, K Furukawa (1992) Analysis of bph operon from the polychlorinated biphenyl-degrading strain of Pseudomonas pseudoalcaligenes KF707. J Biol Chem 267: 4844–4853. Taira K, N Hayase, N Arimura, S Yamashita, T Miyazaki, K Furukawa (1988) Cloning and nucleotide sequence of the 2,3-dihydroxybiphenyl dioxygenase gene from the PCB-degrading strain of Pseudomonas paucimobilis Q1. Biochemistry 27: 3990–3996. Vaillancourt FH, G Labbé, NM Drouin, PD Fortin, LD Eltis (2002) The mechanism-based inactivation of 2,3dihydroxybiphenyl 1,2-dioxygenase by catecholic substrates. J Biol Chem 277: 2019–2027. van den Tweel WJJ, JB Kok, JAM de Bont (1987) Reductive dechlorination of 2,4-dichlorobenzoate to 4-chlorobenzoate and hydrolytic dehalogenation of 4-chloro, 4-bromo, and 4-iodobenzoate by Alcaligenes denitrificans NTB-1. Appl Environ Microbiol 53: 810–815. Wiegel J, X Zhang, Q Wu (1999) Anaerobic dehalogenation of hydroxylated polychlorinated biphenyls by Desulfitobacterium dehalogenans. Appl Environ Microbiol 65: 2217–2221. Wu Q, DL Bedard, J Wiegel (1997a) Effect of incubation temperature on the route of microbial reductive dechlorination of 2,3,4,6-tetrachlorobiphenyl in polychlorinated biphenyl (PCB)-contaminated and PCB-free freshwater sediments. Appl Environ Microbiol 63: 2836–2843. Wu Q, DL Bedard, J Wiegel (1997b) Temperature determines the pattern of anaerobic microbial dechlorination of Arochlor 1260 primed by 2,3,4,6-tetrachlorobiphenyl in Woods Pond sediment. Appl Environ Microbiol 63: 4818–4825. Wu Q, DL Bedard, J Wiegel (1999) 2,6-dibromobiphenyl primes extensive dechlorination of Arochlor 1260 in contaminated sediment at 8–30°C by stimulating growth of PCB-dehalogenating microorganisms. Environ Sci Technol 33: 595–602. Wu Q, JEM Watts, KR Sowers, HD May (2002) Identification of a bacterium that specifically catalyzes the reductive dechlorination of polychlorinated biphenyls with doubly flanked chlorines. Appl Environ Microbiol 68: 807–812. Xun L, E Topp, CS Orser (1992c) Purification and characterization of a tetrachloro-p-hydroquinone reductive dehalogenase from a Flavobacterium sp. J Bacteriol 174: 8003–8007. Yadav JS, JF Quensen, JM Tiedje, CA Reddy (1995) Degradation of polychlorinated biphenyl mixtures (Arochlors 1242, 1254, 1260) by the white rot fungus Phanerochaete chrysosporium as evidenced by congener-specific analysis. Appl Environ Microbiol 61: 2560–2565. Yan D-Z, H Liu, N-Y Zhou (2006) Conversion of Sphingobium chlorophenolicum ATCC 39723 to a hexachlorobenzene degrader by metabolic engineering. Appl Environ Microbiol 72: 2283–2286. Yoshida N, N Takahashi, A Hiraishi (2005) Phylogenetic characterization of polychlorinated-dioxindechlorinating microbial community by use of microcosm studies. Appl Environ Microbiol 71: 4325–4334.

PART 2 HALOGENATED (CHLORINE, BROMINE, AND IODINE) PHENOLS AND ANILINES PHENOLS Aerobic Conditions Chlorinated phenols are used for the impregnation of timber and the production of phenoxyalkanoate herbicides that are degraded by dioxygenation to chlorophenols. Several mechanisms are used for the aerobic degradation of halogenated phenols by bacteria. Those with three or fewer halogen substituents are converted into halogenated catechols by hydroxylases (Beadle and Smith 1982; Perkins et al. 1990) before ring fission (Figure 9.19). This contrasts with the initial reactions for halogenated arenes, 3- and 4-halogenated benzoates, and halogenated amines that are converted by dioxygenation into cis-dihydrodiols before dehydrogenation to halogenated catechols. Therefore, these occupy a central position in the metabolism of all these compounds

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483

OH

OH

OH

Cl Cl

CO 2 H CO 2 H

Cl

Cl

HO 2 C

O

O

Cl

Cl

Cl

HO 2 C

HO 2 C

O

HO 2 C

O

O

CO 2 H

CO 2 H

CO 2 H

Cl

FIGURE 9.19

Aerobic degradation of 2,4-dichlorophenol. OH

NH 2

CO 2 H

Cl

Cl

Cl

H2N

HO 2 C

OH

Cl

OH OH

H

H

Cl

Cl

H

HO

OH

H OH

Cl

OH OH

Cl

FIGURE 9.20 Outline of the biodegradation of 4-chlorophenol, 4-chloroaniline, 4-chlorobenzoate, and chlorobenzene.

(Müller et al. 2004) (Figure 9.20). In contrast, halogenated phenols with three or more substituents are degraded by pathways in which the halogen is removed by various reactions before ring fission: 1. Hydroxylation and dioxygenation are not, however, mutually exclusive, because the toluene dioxygenase from Pseudomonas putida F1 hydroxylates both phenol and 2,5-dichlorophenol with the introduction of only one atom of oxygen (Spain et al. 1989). Subsequent degradation by dioxygenation follows established pathways in which chloride is eliminated from muconic acids after ring fission. 2. Dioxygenation of chlorinated hydroquinones with ring fission has also been observed. The degradation of pentachlorophenol by Sphingomonas chlorophenolica ATCC 39723 proceeds by a series of steps to 2,6-dichlorohydroquinone that undergoes dioxygenation and ring fission to an acyl chloride that is hydrolyzed to 2-chloromaleylacetate (Figure 9.21)

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Environmental Degradation and Transformation of Organic Chemicals OH

Cl

Cl

Cl

CO 2 H COCl

H OH

FIGURE 9.21

Cl

CO 2 H CO 2 H

CO 2 H CO 2 H

CO 2 H CO 2 H

O

O

O

H OH

Degradation of 2,6-dichlorohydroquinone.

(Xun et al. 1999; Ohtsubo et al. 1999). The degradation of 2-chlorohydroquinone, which is produced during the degradation of γ-hexachlorocyclohexane, is carried out analogously by dioxygenation followed by hydrolysis of the acyl chloride to 3-hydroxymuconate, which is mediated by an unusual extradiol fission enzyme encoded by linE (Miyauchi et al. 1999; Endo et al. 2005). In addition to hydrolytic replacement of halogen, reductive displacement has been shown to occur during the degradation of a few aromatic compounds, even under aerobic conditions. For example, both hydrolytic and reductive reactions are involved in the degradation of a number of chlorophenols with three or more substituents. Degradation takes place by a pathway that is initiated by oxidative dechlorination with the formation of hydroquinones followed by formal displacement of chlorine substituents by hydroxyl, or by reduction before ring cleavage. These reactions have been shown to occur in the degradation of a number of polychlorinated phenols, including 2,4,5- and 2,4,6-trichlorophenol and pentachlorophenol, which has attracted particular attention in view of its toxicity. The following examples illustrate initiation by monooxygenation followed by further loss of halogen before ring fission: 1. The degradation of 2,4,6-trichlorophenol has been examined in a number of bacteria. Monooxygenation plays a key role, and 2,6-dichlorobenzoquinone and 6-chlorohydroxyquinol have been recognized as intermediates: a. The monooxygenase that produced 2,6-dichlorohydroquinone as the initial metabolite has been purified and characterized from Azotobacter sp. strain GP1, and is able to accept a number of other polychlorinated phenols (Wieser et al. 1997). b. In Ralstonia eutropha strain JMP134, both monooxygenation to 2,6-dichlorohydroquinone and hydrolysis to 2-chloro-6-hydroxybenzoquinone are accomplished by a single enzyme, and this was confirmed by the incorporation of 18O from both 18O2 and H218O (Xun and Webster 2004). 2-Chloro-6-hydroxybenzoquinone is not, however, a necessary intermediate, since a dioxygenase can bring about ring fission of 2,6-dichlorohydroquinone directly to produce 2-chloromaleylacetate (Figure 9.22a) (Xun et al. 1999; Ohtsubo et al. 1999). Details including the genetics and biochemistry have been given by Louie et al. (2002). c. In Streptomyces rochei strain 303, 2,6-dichlorohydroquinone and 6-chloro-2-hydroxyhydroquinone are produced, and the dioxygenase that brings about intradiol fission of this to 2-chloromaleylacetate has been purified. The NH2-terminal amino acid sequence showed a high degree of similarity to the corresponding enzyme from Azotobacter sp. strain GP1 (Zaborina et al. 1995). 2. The degradation of 2,4,5-trichlorophenoxyacetate by Burkholderia (Pseudomonas) cepacia AC1100 (Haugland et al. 1990) is initiated by a 2-ketoglutarate-dependent dioxygenation to 2,4,5-trichlorophenol, which is degraded by initial monooxygenation and reduction to 2,5-dichlorohydroquinone and 5-chloro-2-hydroxy-hydroquinone (Daubaras et al. 1995; Gisis and Xun 2003). This is then dechlorinated to 2-hydroxyhydroquinone before ring fission to maleylacetate (Figure 9.22b) (Daubaras et al. 1996; Latus et al. 1995; Zaborina

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Substituted Carbocyclic Aromatic Compounds OH Cl

485

OH Cl

Cl

Cl

Cl

(a)

CO 2 H COCl

H Cl

OH

OH

H

CO 2 H CO 2 H

H

OH

OH Cl

Cl

OH

OH

OH Cl

OH

OH

(b) Cl

Cl Cl

Cl OH

OH

OH H

CO 2 H CO 2 H

CO 2 H CO 2 H

H O

FIGURE 9.22

O

Degradation of (a) 2,4,6-trichlorophenol and (b) 2,4,5-trichlorophenol.

et al. 1998). This strain also brings about para-hydroxylation of dichlorophenols, independent of the existence of a chlorine substituent (Tomasi et al. 1995). 3. The degradation of pentachlorophenol has been studied in both Gram-positive strains of Mycobacterium sp. (Suzuki 1983) and Mycobacterium (Rhodococcus) chlorophenolicum (Apajalahti and Salkinoja-Salonen 1987b), and in the Gram-negative Sphingobium (Sphingomonas) chlorophenolicum [Flavobacterium sp.] (Steiert and Crawford 1986) and strain RA2 (Miethling and Karlson 1996). In the application of such strains to bioremediation of contaminated sites, a number of additional factors should be taken into consideration including their tolerance to high concentrations of pentachlorophenol and their longevity in the environment. a. The pathway was delineated for Mycobacterium chlorophenolicum (Rhodococcus chlorophenolicus) strain PCP-1, and is initiated as for 2,4,6- and 2,4,5-trichlorophenols by monooxygenation with the formation of tetrachlorohydroquinone. Subsequent steps involve hydrolytic dechlorination followed by three reductive dechlorinations to 1,2,4-trihydroxybenzene, which undergo ring fission (Apajalahti and SalkinojaSalonen 1987a,b; Uotila et al. 1995) (Figure 9.23a). In Mycobacterium fortuitum strain CG-2, the monooxygenase that initiates degradation is membrane associated, whereas the enzyme that carried out hydroxylation and reductive dechlorination to 1,2,4-trihydroxybenzene was soluble (Uotila et al. 1992). b. Degradation by Sphingobium chlorophenolicum ATCC 39723 is initiated by the formation of tetrachlorohydroquinone via the corresponding tetrachlorobenzoquinone (Dai et al. 2003), and the pathway for its degradation has been described (Cai and Xun 2002). Sequential reductive dechlorinations to 2,3,6-trichlorohydroquinone and 2,6dichlorohydroquinone are mediated by a glutathione S-transferase system (Orser et al. 1993). Although a putative chlorohydrolase has been implicated in the degradation of 2,6-dichlorohydroquinone (Lee and Xun 1997; Chanama and Crawford 1997), the substrate can be oxidized with fission of the ring to 2-chloromaleylacetate before reductive dechlorination to maleylacetate (Kaschabek and Reineke 1995) (Figure 9.23b).

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Environmental Degradation and Transformation of Organic Chemicals OH

OH Cl

Cl

Cl

Cl

Cl

OH Cl

Cl

Cl

Cl

OH OH

OH

HO 2 C

H

(a) Cl Cl

OH

OH

OH

Cl

H

OH

OH

OH

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

CO 2 H

OH

Cl

Cl

Cl

(b)

Cl Cl

CO 2 H CO 2 H

H

Cl

CO 2 H CO 2 H

CO 2 H CO 2 H

H OH

FIGURE 9.23

OH

OH

OH

O

O

Pathways for biodegradation of pentachlorophenol.

Comparable pathways may bring about dehalogenation of phenols with other para-halogen substituents: a. The pentachlorophenol monooxygenase from Sphingobium chlorophenolicum (Flavobacterium sp.) ATCC 39723 is able to hydroxylate a range of other para-substituted phenols with elimination of, for example, chloride, bromide, iodide, cyanide, and nitrite from the 4-position (Xun et al. 1992a,b,c). b. Loss of fluoride from pentafluorophenol and bromide from pentabromophenol is catalyzed by the monooxygenase from Mycobacterium fortuitum strain CG-2 (Uotila et al. 1992). Fungi and Yeasts Although these have been less exhaustively investigated than their bacterial counterparts, the results of these investigations have revealed a number of significant features: 1. The NIH shift (Daly et al. 1972) with translocation of chlorine has been demonstrated during the biotransformation of 2,4-dichlorophenoxyacetate by Aspergillus niger (Figure 9.24) (Faulkner and Woodcock 1965). The NIH shift is not restricted to fungi since it has also been demonstrated with protons—though less frequently with other substituents—in prokaryotes. 2. The degradation of chlorinated phenols has been examined with the white-rot basidiomycete Phanerochaete chrysosporium under conditions of nitrogen limitation, and apparently involves both lignin peroxidase and manganese-dependent peroxidase activities (Valli and Gold 1991). For 2,4-dichlorophenol, the reaction involves a series of oxidations and reductions (Figure 9.25) that are entirely different from the sequence that is employed by bacteria. Essentially similar reactions are involved in the degradation of 2,4,5-trichlorophenol, and this is accomplished more rapidly—possibly due to less interference from

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Substituted Carbocyclic Aromatic Compounds

487

O−CH 2−CO2H

O−CH 2−CO2H

O−CH 2−CO2H

Cl

Cl

Cl

HO

Cl

Cl

FIGURE 9.24

Cl

OH

NIH shift during the metabolism of 2,4-dichlorophenoxyacetate by Aspergillus niger.

OH

OH

O Cl

Cl

OCH 3 Cl

OCH 3 Cl

O

O Cl

OH

O

OCH 3

OCH 3

O OH

OCH 3

OH OH

OH

CO 2 H CH 2

HO

HO OCH 3

HO O

CO 2 H

OH

FIGURE 9.25 Biodegradation of 2,4-dichlorophenol by Phanerochaete chrysosporium.

polymerization reactions (Joshi and Gold 1993). In addition, 2-chloro-1,4-dimethoxybenzene is produced from 2,4-dichlorophenol, and 2,5-dichloro-1,4-dimethoxybenzene from 2,4,5-trichlorophenol that also produced small amounts of a dimeric product that was tentatively identified as 2,2′-dihydroxy-3,3′,5,5′,6,6′-hexachlorobiphenyl. The degradation of 2,4,6-trichlorophenol by P. chrysosporium is analogous to that of 2,4-dichlorophenol, and forms 2-chloro-1,4-dihydroxybenzene that is converted by alternative pathways to 1,2,4-trihydroxybenzene before ring cleavage (Reddy et al. 1998). Parenthetically, a formally comparable sequence of reactions is used for the degradation of 2,4-dinitrotoluene (Valli et al. 1992). The degradation of pentachlorophenol is initiated by formation of tetrachloro-1,4-benzoquinone, and is followed by a series of reductive dechlorinations to produce 1,2,4-trihydroxybenzene that undergoes ring fission (Reddy and Gold 2000). 3. The yeast Candida maltosa is capable of assimilating phenol, and phenol-grown cells were able to hydroxylate 2-chlorophenol to 3-chlorocatechol and bring about ring fission to cis,cis-2-chloromuconate. 3-Chlorophenol and 4-chlorophenol produced 4-chlorocatechol that was subsequently hydroxylated to 5-chloropyrogallol, which underwent ring fission to 4-carboxymethylenebut-2-en-4-olide (Polnisch et al. 1992). Anaerobic Conditions The persistence of halogenated phenols and anilines in anaerobic environments is determined by the activity of anaerobic dehalogenating bacteria. Extensive effort has therefore been devoted to isolating the relevant organisms and an increasing number of strains have been obtained in pure culture. They display different specificities for the position of the halogen; several bring about dechlorination specifically at positions ortho to the chlorine, while some of them can use the aromatic substrate as

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an electron acceptor. The role of reductive dechlorination in energy metabolism (dehalorespiration) has been reviewed (Holliger et al. 1999): 1. Cells of Desulfomonile tiedjei induced by growth with 3-chlorobenzoate were able to partially dechlorinate polychlorinated phenols, specifically at the 3-position, whereas the monochlorophenols were apparently resistant to dechlorination (Mohn and Kennedy 1992). Details of the reductase have already been described in Part 1 of this chapter. 2. A sulfite-reducing organism Desulfitobacterium dehalogenans, which is unable to carry out dissimilatory reduction of sulfate, dechlorinates 3-chloro-4-hydroxyphenylacetate to 4-hydroxyphenylacetate as a terminal metabolite during growth with pyruvate (Utkin et al. 1994). The specificity of partial dechlorination of ortho-chlorinated phenols has been examined (Utkin et al. 1995). The reductive halogenase was characterized, and electron paramagnetic resonance (EPR) analysis showed the presence of one [4Fe–4S] cluster and one cobalamin per monomer (van de Pas et al. 1999). This strain is also able to completely dehalogenate biphenyls with ortho chlorine and hydroxyl substituents, such as 3,3′, 5,5′-tetrachloro-4,4′dihydroxybiphenyl, which is a PCB metabolite from higher organisms (Wiegel et al. 1999). This preference for the presence of an ortho-hydroxyl group has been explored in Desulfitobacterium chlororespirans and was attributed to its necessity for binding to the substrate (Krasotkina et al. 2001). 3. D. hafniense (frappieri) strain PCP-1 was isolated from a methanogenic consortium and could dechlorinate pentachlorophenol to 3-chlorophenol via 2,3,4,5-tetra-, 3,4,5-tri-, and 3,5-dichlorophenol (Bouchard et al. 1996). In addition, it is able to dechlorinate a wide range of polychlorinated aromatic substrates including phenols, catechols, anilines, pentachloronitrobenzenes, and pentachloropyridine (Dennie et al. 1998). A 3,5-dichlorophenol reductive dehalogenase has been characterized from this strain, and is able to dechlorinate a number of phenols at the meta and para positions (Thibodeau et al. 2004). The organism assigned to D. hafniense strain DCB-2 is able to dechlorinate 3-chloro-4-hydroxyphenylacetate to 4-hydroxyphenylacetate and 2,4,6-trichlorophenol to 4-chlorophenol. The membrane-bound dehalogenase has been characterized and contains per mol of subunit 0.7 mol corrinoid, 12 mol iron, and 13 mol acid-labile sulfur (Christiansen et al. 1998). 4. A facultatively anaerobic organism designated Anaeromyxobacter dehalogenans (Sanford et al. 2002) was capable of dechlorinating ortho-chlorinated phenols using acetate as electron donor—2-chlorophenol was reduced to phenol and 2,6-dichlorophenol to 2-chlorophenol (Cole et al. 1994). A strain of Desulfovibrio dechloracetivorans was also able to couple the dechlorination of ortho-substituted chlorophenols to the oxidation of acetate, fumarate, lactate, and propionate (Sun et al. 2000). 5. A spore-forming strain of Desulfitobacterium chlororespirans was able to couple the dechlorination of 3-chloro-4-hydroxybenzoate to the oxidation of lactate to acetate, pyruvate, or formate (Sanford et al. 1996). Whereas 2,4,6-trichlorophenol and 2,4,6-tribromophenol supported growth with the production of 4-chlorophenol and 4-bromophenol, neither 2-bromophenol nor 2-iodophenol was able to do so. The membrane-bound dehalogenase contains cobalamin, iron, and acid-labile sulfur, and is apparently specific for ortho-substituted phenols (Krasotkina et al. 2001). 6. Desulfitobacterium chlororespirans can use ortho-substituted phenols as electron acceptors for anaerobic growth, and is able to debrominate 2,6-dibromo-4-cyanophenol (Bromoxynil) and 2,6-dibromo-4-carboxyphenol. In contrast, 2,6-diiodo-4-cyanophenol (Ioxynil) was deiodinated only in the presence of 3-chloro-4-hydroxybenzoate (Cupples et al. 2005). 7. Desulfovibrio strain TBP-1 grown with lactate and sulfate as electron acceptor is able to debrominate 2-bromo-, 2,4-dibromo, 2,6-dibromo-, and 2,4,6-tribromophenol (Boyle et al. 1999).

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Substituted Carbocyclic Aromatic Compounds

489 Cl

Cl Cl

Cl

(a)

Cl

Cl

Cl

Cl

Cl

Cl Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl Cl

Cl

Cl

Cl

N H2

N H2

N H2

Cl

Cl

Cl

(b) Cl

Cl

Cl

Cl

Cl

Cl

Cl N H2

N H2 Cl

N H2

N H2 Cl

; Cl

Cl Cl

FIGURE 9.26

Cl

Cl

Partial anaerobic dechlorination of (a) hexachlorobenzene and (b) 2,3,4,5-tetrachloroaniline.

In addition to these investigations using pure strains, some have used enrichment cultures in various experimental designs that are important to appreciate in interpreting the results. Complete dechlorination of polyhalogenated compounds under anaerobic conditions has seldom been observed, and most commonly only partial dehalogenation occurs: all these reactions are therefore strictly biotransformations. Illustrative examples include hexachlorobenzene (Figure 9.26a) (Fathepure et al. 1988) and tetrachloroaniline (Figure 9.26b) (Kuhn et al. 1990), although pentachlorophenol can be completely dechlorinated and mineralized (Mikesell and Boyd 1986). The flame-retardant tetrabromobisphenol A was debrominated under both methanogenic and sulfidogenic conditions (Ronen and Abeliovich 2000; Voordeckers et al. 2002).

ANILINES Aerobic Conditions Chlorinated anilines are produced by the hydrolysis of a range of acetanilide, urea, and carbamate herbicides, and are therefore widely distributed in agricultural soils. Mechanisms for their loss

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include not only biodegradation but also the formation of stable complexes with humic material. The range of halogenated anilines that have been examined includes the following: a. 2-, 3-, and 4-chloroanilines, and 4-fluoroaniline by a Moraxella sp. strain G (Zeyer et al. 1985), and 3- and 4-chloroaniline by a strain of Pseudomonas acidovorans (Loidl et al. 1990) b. 2-Methylaniline and 4-chloro-2-methylaniline by Rhodococcus rhodochrous strain CTM (Fuchs et al. 1991) c. 3-Chloro-4-methylaniline by Burkholderia (Pseudomonas) cepacia strain CMA1 (Stockinger et al. 1992) d. 3,4-Dichloroaniline by Variovorax sp. strain WDL1 (Dejonghe et al. 2003) Degradation of chloroanilines is initiated by dioxygenation with the elimination of NH4+, followed by ring fission of the resulting chlorocatechols. The pathway for 4-chloroaniline involves oxidative deamination to 4-chlorocatechol and intradiol ring fission to the same intermediates that are involved in the degradation of halogenated catechols (Zeyer et al. 1985). From the broad specificity of aniline oxygenase, it is plausible to assume that the pathway is the same for all 4-halogenated anilines. The degradation of 3-chloroaniline has been examined in strains of Comamonas testosteroni and Delftia (Pseudomonas) acidovorans, and is initiated by dioxygenation with the production of 4-chlorocatechol that could undergo either intradiol or extradiol ring fission (Boon et al. 2001). The herbicide linuron, which is a urea derivative of 3,4-dichloroaniline, can be degraded by Variovorax sp. strain WDL1 to 3,4-dichloroaniline before fission of 3,4- or 4,5-dichlorocatechol by an unresolved pathway (Dejonghe et al. 2003). There are, however, a number of complicating issues that determine the degradability of chloroanilines: 1. The degradation of aniline may be induced by aniline, although both 3- and 4- chloroaniline, which are poor substrates, were able to induce the enzymes for aniline degradation in a strain of Pseudomonas sp. (Konopka et al. 1989). This strain was able to degrade aniline in the presence of readily degradable substrates such as lactate. 2. The degradation of 3- and 4-chloroaniline may require the presence of either aniline or glucose (references in Zeyer et al. 1985), while the metabolism of methylanilines required the addition of ethanol as additional carbon source (Fuchs et al. 1991). 3. The degradation of 3-chloro-4-methylaniline by Pseudomonas cepacia strain CMA1 involved ring fission of 3-chloro-4-methylcatechol by an intradiol enzyme (Stockinger et al. 1992). 4. The gene tdnQ that encodes the oxidative deamination of aniline is carried on a plasmid in strains of Comamomas testosteroni and Delftia acidovorans. However, it has not been resolved whether this enzyme is involved in the degradation of 3-chloroaniline, because the gene is transcribed in the presence of aniline, but not when only 3-chloroaniline is present (Boon et al. 2001). Anaerobic Conditions The partial dechlorination of chlorinated anilines has been examined in anaerobic slurries, which has already been noted (Kuhn et al. 1990), while D. hafniense (frappieri) strain PCP-1 is able to dechlorinate a wide range of polychlorinated aromatic substrates including phenols, catechols, anilines, pentachloronitrobenzene, and pentachloropyridine (Dennie et al. 1998). Both pentachloroaniline and 2,3,5,6-tetrachloroaniline were partially dechlorinated to dichloroanilines, although the orientation of the substituents was unresoved. The kinetics of dechlorination of pentachloroaniline have been described (Tas et al. 2006).

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491

REFERENCES Apajalahti JHA, MS Salkinoja-Salonen (1987a) Dechlorination and para-hydroxylation of polychlorinated phenols by Rhodococcus chlorophenolicus. J Bacteriol 169: 675–681. Apajalahti JHA, MS Salkinoja-Salonen (1987b) Complete dechlorination of tetrachlorohydroquinone by cell extracts of pentachlorophenol-induced Rhodococcus chlorophenolicus. J Bacteriol 169: 5125–5130. Beadle CA, ARW Smith (1982) The purification and properties of 2,4-dichlorophenol hydroxylase from a strain of Acinetobacter sp. Eur J Biochem 123: 323–332. Boon N, J Goris, P de Vos, W Verstraete, EM Top (2001) Genetic diversity among 3-chloroaniline- and anilinedegrading strains of Comomonadaceae. Appl Environ Microbiol 67: 1107–1115. Bouchard B, R Beaudet, R Villemur, G MscSween, F Lépine, J-G Bisaillon (1996) Isolation and characterization of Desulfitobacterium frappieri sp. nov., an anaerobic bacterium which reductively dechlorinates pentachlorophenol to 3-chlorophenol. Int J Syst Bacteriol 46: 1010–1015. Boyle AW, CD Phelps, LY Young (1999) Isolation from estuarine sediments of a Desulfovibrio strain which can grow on lactate coupled to the reductive dehalogenation of 2,4,6-tribromophenol. Appl Environ Microbiol 65: 1133–1140. Cai M, L Xun (2002) Organization and regulation of pentachlorophenol-degrading genes in Sphingobium chlorophenolicum ATCC 39723. J Bacteriol 184: 4672–4680. Chanama S, RL Crawford (1997) Mutational analysis of pcpA and its role in pentachlorophenol degradation by Sphingomonas (Flavobacterium) chlorophenolica ATCC39723. Appl Environ Microbiol 63: 4833–4838. Christiansen N, BK Ahring, G Wohlfarth, G Diekert (1998) Purification and characterization of the 3-chloro-4hydroxyphenylacetate reductive dehalogenase of Desulfitobacterium hafniense. FEBS Lett 436: 159–162. Cole JR, AL Cascarelli, WW Mohn, JM Tiedje (1994) Isolation and characterization of a novel bacterium growing via reductive dehalogenation of 2-chlorophenol. Appl Environ Microbiol 60: 3536–3542. Cupples AM, RA Sanford, GK Sims (2005) Dehalogenation of the herbicides Bromoxynil (3,5-dibromo4-hydroxybenzonitrile) and Ioxynil (3,5-diiodo-4-hydroxybenzonitrile) by Desulfitobacterium chlororespirans. Appl Environ Microbiol 71: 3741–3746. Dai M, JB Rogers, JR Warner, SD Copley (2003) A previously unrecognized step in pentachlorophenol degradation in Sphingobium chlorophenolicum is catalyzed by tetrachlorobenzoquinone reductase (PcpD). J Bacteriol 185: 302–310. Daly J, DM Jerina, B Witkop (1972) Arene oxides and the NIH shift: the metabolism, toxicity and carcinogenicity of aromatic compounds. Experientia 28: 1129–1149. Daubaras DL, CD Hershberger, K Kitano, AM Chakrabarty (1995) Sequence analysis of a gene cluster involved in metabolism of 2,4,5-trichlorophenoxyacetic acid by Burkholderia cepacia AC1100. Appl Environ Microbiol 61: 1279–1289. Daubaras DL, K Saido, AM Chakrabarty (1996) Purification of hydroxyquinol 1,2-dioxygenase and maleylacetate reductase: the lower pathway of 2,4,5-trichlorophenoxyacetic acid metabolism by Burkholderia cepacia AC1100. Appl Environ Microbiol 62: 4276–4279. Dejonghe W, E Berteloot, J Goris, N Boon, K Crul, S Maertens, M Höfte, P de Vos, W Verstraete, EM Top (2003) Synergistic degradation of linuron by a bacterial consortium and isolation of a single linurondegrading Variovorax strain. Appl Environ Microbiol 69: 1532–1542. Dennie D, I Gladu, F Lépine, R Villemur, J-G Bisaillon, R Beaudet (1998) Spectrum of the reductive dehalogenation activity of Desulfitobacterium frappieri PCP-1. Appl Environ Microbiol 64: 4603–4606. Endo R, M Kamakura, K Miyauchi, M Fukuda, Y Ohtsubo, M Tsuda, Y Nagata (2005) Identification and characterization of genes involved in the downstream degradation pathway of γ-hexachlorocyclohexane in Sphingomonas paucimobilis UT26. J Bacteriol 187: 847–853. Fathepure BZ, JM Tiedje, SA Boyd (1988) Reductive dechlorination of hexachlorobenzene to tri- and dichlorobenzenes in anaerobic sewage sludge. Appl Environ Microbiol 54: 327–330. Faulkner JK, D Woodcock (1965) Fungal detoxication. Part VII. Metabolism of 2,4-dichlorophenoxyacetic and 4-chloro-2-methylphenoxyacetic acids by Aspergillus niger. J Chem Soc 1187–1191. Fuchs K, A Schreiner, F Lingens (1991) Degradation of 2-methylaniline and chlorinated isomers of 2-methylaniline by Rhodococcus rhodochrous strain CTM. J Gen Microbiol 137: 2033–2039. Gisi MR, L Xun (2003) Characterization of chlorophenol 4-monooxygenase (TftD) and NADH: flavin adenine dinucleotide oxidoreductase (TftC) of Burkholöderia cepacia AC100. J Bacteriol 185: 2786–2792. Haugland RA, DJ Schlemm, RP Lyons, PR Sferra, AM Chakrabarty (1990) Degradation of the chlorinated phenoxyacetate herbicides 2,4-dichlorophenoxyacetic acid and 2,4,5-trichlorophenoxyacetic acid by pure and mixed bacterial cultures. Appl Environ Microbiol 56: 1357–1362.

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Holliger C, G Wohlfarth, G Diekert (1999) Reductive dechlorination in the energy metabolism of anaerobic bacteria. FEMS Microbiol Revs 22: 383–398. Joshi DK, MH Gold (1993) Degradation of 2,4,5-trichlorophenol by the lignin-degrading basidiomycete Phanerochaete chrysosporium. Appl Environ Microbiol 59: 1779–1785. Kaschabek SR, W Reineke (1995) Maleylacetate reductase of Pseudomonas sp. strain B13: specificity of substrate conversion and halide elimination. J Bacteriol 177: 320–325. Konopka A, D Knoght, RF Turco (1989) Characterization of a Pseudomonas sp. capable of aniline degradation in the presence of secondary carbon sources. Appl Environ Microbiol 55: 385–389. Krasotkina J, T Walters, KA Maruya, SW Ragsdale (2001) Characterization of the B12- and iron–sulfurcontaining reductive dehalogenase from Desulfitobacterium chlororespirans. J Biol Chem 276: 40991–40997. Kuhn EP, GT Townsend, JM Suflita (1990) Effect of sulfate and organic carbon supplements on reductive dehalogenation of chloroanilines in anaerobic aquifer slurries. Appl Environ Microbiol 56: 2630–2637. Latus M, H-J Seitz, J Eberspächer, F Lingens (1995) Purification and characterization of hydroxyquinol 1,2dioxygenase from Azotobacter sp. strain GP1. Appl Environ Microbiol 61: 2453–2460. Lee J-Y, L Xun (1997) Purification and characterization of 2,6-dichloro-p-hydroquinone chlorohydrolase from Flavobacterium sp. strain ATCC 39723. J Bacteriol 179: 1521–1524. Loidl M, C Hinteregger, G Ditzelmüller, A Ferschl, F Streichsbier (1990) Degradation of aniline and monochlorinated anilines by soil-borne Pseudomonas acidovorans strains. Arch Microbiol 155: 56–61. Louie TM, CM Webster, L Xun (2002) Genetic and biochemical characterization of a 2,4,6-trichlorophenol degradation pathway in Ralstonia eutropha JMP134. J Bacteriol 184: 3492–3500. Miethling R, Karlson U (1996) Accelerated mineralization of pentachlorophenol in soil upon inoculation with Mycobacterium chlorophenolicum PCP1 and Sphingomonas chlorophenolica RA2. Appl Environ Microbiol 62: 4361–4366. Mikesell MD, SA Boyd (1986) Complete reductive dechlorination and mineralization of pentachlorophenol by anaerobic microorganisms. Appl Environ Microbiol 52: 861–865. Miyauchi K, Y Adachi, Y Nagata, M Takagi (1999) Cloning and sequencing of a novel meta-cleavage dioxygenase gene whose product is involved in degradation of γ-hexachlorocyclohexane in Sphingomonas paucimobilis. J Bacteriol 181: 6712–6719. Mohn WW, KJ Kennedy (1992) Reductive dehalogenation of chlorophenols by Desulfomonile tiedjei DCB-1. Appl Environ Microbiol 58: 1367–1370. Müller TA, SM Byrde, C Werlen, JR van der Meer, H-P Kohler (2004) Genetic analysis of phenoxyalkanoic acid degradation in Sphingomonas herbicidovorans MH. Appl Environ Microbiol 70: 6066–6075. Ohtsubo Y, K Miyauchi, K Kanda, T Hatta, H Kiyohara, T Senda, Y Nagata, Y Mitsui, M Takagi (1999) PchA, which is involved in the degradation of pentachlorophenol in Sphingomonas chlorophenolica ATCC 39723, is a novel type of ring-cleavage dioxygenase. FEBS Letters 459: 395–398. Orser CS, J Dutton, C Lange, P Jablonsli, L Xun, M Hargis (1993) Characterization of a Flavobacterium glutathione S-transferase gene involved in reductive dechlorination. J Bacteriol 175: 2640–2644. Perkins EJ, MP Gordon, O Caceres, PF Lurquin (1990) Organization and sequence analysis of the 2,4dichlorophenol hydroxylase and dichlocatechol oxidative operons of plasmid pJP4. J Bacteriol 172: 2351–2359. Polnisch E, H Kneifel, H Franzke, KL Hofmann (1992) Degradation and dehalogenation of monochlorophenols by the phenol-assimilating yeast Candida maltosa. Biodegradation 2: 193–199. Reddy GVB, MDS Gelpke, MH Gold (1998) Degradation of 2,4,6-trichlorophenol by Phanerochaete chrysosporium: involvement of reductive dechlorination. J Bacteriol 180: 5159–5164. Reddy GVB, MH Gold (2000) Degradation of pentachlorophenol by Phanerochaete chrysosporium: intermediates and reactions involved. Microbiology (UK) 146: 405–413. Ronen Z, A Abeliovich (2000) Anaerobic–aerobic process for microbial degradation of tetrabromobisphenol-A. Appl Environ Microbiol 66: 2372–2377. Sanford RA, JR Cole, FE Löffler, JM Tiedje (1996) Characterization of Desulfitobacterium chlororespirans sp. nov., which grows by coupling the oxidation of lactate to the reductive dechlorination of 3-chloro-4hydroxybenzoate. Appl Environ Microbiol 62: 3800–3808. Sanford RA, JR Cole, JM Tiedje (2002) Characterization and description of Anaeromyxobacter dehalogenans gen., nov., sp. nov., an aryl-halorespiring facultative anaerobic myxobacterium. Appl Environ Microbiol 68: 893–900. Spain JC, GJ Zylstra, CK Blake, DT Gibson (1989) Monohydroxylation of phenol and 2,5-dichlorophenol by toluene dioxygenase in Pseudomonas putida F1. Appl Environ Microbiol 55: 2648–2652.

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Steiert JG, RL Crawford (1986) Catabolism of pentachlorophenol by a Flavobacterium sp. Biochem Biophys Res Commun 141: 1421–1427. Stockinger J, C Hinteregger, M Loidl, A Ferschl, F Streichsbier (1992) Mineralizarion of 3-chloro-4-methylamiline via an ortho-cleavage pathway by Pseudomonas cepacia strain CMA1. Appl Microbiol Biotechnol 38: 421–428. Sun B, JR Cole, RA Sanford, JM Tiedje (2000) Isolation and characterization of Desulfobvibrio dechloracetivorans sp. nov., a marine dechlorinating bacterium growing by coupling the oxidation of acetate to the reductive dechlorination of 2-chlorophenol. Appl Environ Microbiol 66: 2408–2413. Suzuki T (1983) Methylation and hydroxylation of pentachlorophenol by Mycobacterium sp. isolated from soil. J Pesticide Sci 8: 419–428. Tas DO, IN Thomson, FE Löffler, SG Pavlostathis (2006) Kinetics of the microbial reductive dechlorination of pentachloroaniline. Environ Sci Technol 40: 4467–4472. Thibodeau J, A Gauthier, M Duguay, R Villemur, F Lépine, P Juteau, R Beaudet (2004) Purification, cloning, and sequencing of a 3,5-dichlorophenol reductive dehalogenase from Desulfitobacterium frappieri PCP-1. Appl Environ Microbiol 70: 4532–4537. Tomasi I, I Artaud, Y Bertheau, D Mansuy (1995) Metabolism of polychlorinated phenols by Pseudomonas cepacia AC1100: determination of the first two steps and specific inhibitory effect of methimazole. J Bacteriol 177: 307–311. Uotila JS, VH Kitunen, T Coote, T Saastamoinen, MS Salkinoja-Salonen, JHA Apajalahtri (1995) Metabolism of halohydroquinones in Rhodococcus chlorophenolicus PCP-1. Biodeg 6: 119–126. Uotila JS, VH Kitunen, T Saastamoinen, T Coote, MM Häggblom, MS Salkinoja-Salonen (1992). Characterization of aromatic dehalogenases of Mycobacterium fortuitum CG-2. J Bacteriol 174: 5669–5675. Utkin I, C Woese, J Wiegel (1994) Isolation and characterization of Desulfitobacterium dehalogenans gen. nov., sp. nov., an anaerobic bacterium which reductively dechlorinates chlorophenolic compounds. Int J Syst Bacteriol 44: 612–619. Utkin I, DD Dalton, J Wiegel (1995) Specificity of reductive dehalogenation of substituted orthochlorophenols by Desulfitobacterium dehalogenans JW/IU-DC1. Appl Environ Microbiol 61: 346–351. Valli K, BJ Brock, DK Joshi, MH Gold (1992) Degradation of 2,4-dinitrotoluene by the lignin-degrading fungus Phanerochaete chrysosporium. Appl Environ Microbiol 58: 221–228. Valli K, MH Gold (1991) Degradation of 2,4-dichlorophenol by the lignin-degrading fungus Phanerochaete chrysosporium. J Bacteriol 173: 345–352. van de Pas BA, H Smidt, WR Hagen, J van der Oost, G Schraa, AJM Stams, WM de Vos (1999) Purification and molecular characterization of ortho-chlorophenol reductive dehalogenase, a key enzyme of halorespiration in Desulfitobacterium dehalogenans. J Biol Chem 274: 20287–20292. Voordeckers JW, DE Fennell, K Jones, MM Häggblom (2002) Anaerobic biotransformation of tetrabromobisphenol-A, tetrachlorobisphenol-A, and bisphenol A in estuarine sediments. Environ Sci Technol 36: 696–701. Wiegel J, X Zhang, Q Wu (1999) Anaerobic dehalogenation of hydroxylated polychlorinated biphenyls by Desulfitobacterium dehalogenans. Appl Environ Microbiol 65: 2217–2221. Wieser M, B Wagner, J Eberspächer, F Lingens (1997) Purification and characterization of 2,4,6-trichlorophenol-4-monooxygenase, a dehalogenating enzyme from Azotobacter sp. strain GP1. J Bacteriol 179: 202–208. Xun L, CM Webster (2004) A monooxygenase catalyzes sequential dechlorination of 2,4,6-trichlorophenol by oxidative and hydrolytic reactions. J Biol Chem 279: 6696–6700. Xun L, E Topp, CS Orser (1992a) Confi rmation of oxidative dehalogenation of pentachlorophenol by a Flavobacterium pentachlorophenol hydroxylase. J Bacteriol 174: 5745–5747. Xun L, E Topp, CS Orser (1992b) Diverse substrate range of a Flavobacterium pentachlorophenol hydroxylase and reaction stoichiometries. J Bacteriol 174: 2898–2902. Xun L, E Topp, CS Orser (1992c) Purification and characterization of a tetrachloro-p-hydroquinone reductive dehalogenase from a Flavobacterium sp. J Bacteriol 174: 8003–8007. Xun L, J Bohuslavek, M Vai (1999) Characterization of 2,6-dichloro-p-hydroquinone 1,2-dioxygenase (PcpA) of Sphingomonas chlorophenolica ATCC 39723. Biochem Biophys Res Commun 266: 322–325. Zaborina O, DL Daubaras, A Zago, Y Xun, K Saido, T Klem, D Nikolic, AM Chakrabarty (1998) Novel pathway for conversion of chlorohydroxyquinol to maleylacetate in Burkholderia cepacia AC1100. J Bacteriol 180: 4667–4675.

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Zaborina O, M Latus, J Eberspächer, LA Golovleva, F Lingens (1995) Purification and characterization of 6-chlorohydroquinol 1,2-dioxygenase from Streptomyces rochei 303: comparison with an analogous enzyme from Azotobacter sp. strain GP1. J Bacteriol 177: 229–234. Zeyer J, A Wasserfallen, KN Timmis (1985) Microbial mineralization of ring-substituted anilines through an ortho-cleavage pathway. Appl Environ Microbiol 50: 447–453.

PART 3 FLUORINATED HYDROCARBONS, CARBOXYLATES, PHENOLS, AND ANILINES The basic mechanisms for the aerobic degradation of carbocyclic aromatic compounds are well established and initial dioxygenation is now recognized as one of the fundamental mechanisms for activating aromatic rings to aerobic degradation. Many of the mechanisms that operate for chlorinated and brominated arenes apply with relatively minor modification to their fluorinated analogs. Further details are given in a review by Neilson and Allard (2002).

FLUORINATED AROMATIC HYDROCARBONS Aerobic Conditions The transformation of a series of 3-substituted fluorobenzenes has been examined in Pseudomonas strain T-2 in which toluene 2,3-dioxygenase was synthesized (Renganathan 1989). Dioxygenation took place for all of them, and was sometimes accompanied by concomitant loss of fluorine (Figure 9.27). Degradation of fluorobenzene by strain F11 within the order Rhizobiales took place by dioxygenation at C3 and C4 to produce 4-fluorocatechol, or at C1 and C2 to produce catechol. 4-Fluorocatechol was degraded by intradiol fission to 3-ketoadipate

Cl

CH 3

CH 3

OH

HO (23%) HO Cl

Cl

(100%)

F

OH

F CH 3 (3%)

F

HO

F

OH CF 3

(67%) OH

OH Cl

CH 3 F

OH (74%) OH F

(HO)2

(33%) F

F OH (90%) F

OH

FIGURE 9.27 Transformation of fluorinated substrates by toluene 2,3-dioxygenase and yields (%). (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.)

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(Carvalho et al. 2006). The biphenyl dioxygenase BphA in Burkholderia sp. strain LB400, which is able to degrade a wide range of PCBs, was able to transform a number of 2,2′-substituted biphenyls. These included 2,2′-difluorobiphenyl from which a fluorinated 2,3-dihydrodiol was initially formed, and by loss of fluoride this produced 2′-hydroxy-2,3-dihydroxybiphenyl (Seeger et al. 2001). A strain of Sphingomonas strain SS33 adapted to the growth of 4,4′-difluorobiphenyl ether produced the transient intermediates 4-fluorophenol and 4-fluorocatechol before loss of fluoride (Schmidt et al. 1993). These may plausibly occur as a result of the activities of phenol hydroxylase and catechol 1,2-dioxygenase since both these enzymes were found in the parent strain SS33 (Schmidt et al. 1992). Metabolism by Yeasts and Fungi Whereas the metabolism of aromatic hydrocarbons takes place by dioxygenation, their biotransformation by yeasts and fungi is normally initiated by monooxygenation to the epoxide followed by hydrolysis to the trans-dihydrodiols. Phenols may subsequently be formed either by elimination or by nonenzymatic rearrangement of the epoxide: 1. The transformation of 4-fluorobiphenyl by the ectomycorrhizal fungus Tylospora fibrilosa was studied by 19F NMR (Green et al. 1999) and the principal products were 4-fluorobiphen-3′-ol and 4-fluorobiphen-4′-ol. 2. Cunninghamella elegans metabolized 1-fluoronaphthalene to a number of products whose synthesis was clearly initiated by the formation of epoxides and trans-dihydrodiols (Figure 9.28) (Cerniglia et al. 1984). This illustrates the apparent indifference of the monooxygenase to the presence of fluorine atoms. Anaerobic Denitrifying Conditions Some strains originally designated as species of Pseudomonas have now been assigned to other genera such as Thauera or Azoarcus (Anders et al. 1995). There are two major pathways for the anaerobic degradation of toluene, both of which involve formation of benzoate as an intermediate followed by reduction of the benzene ring. One pathway is initiated by reaction with fumarate to produce benzylsuccinate, whereas the other involves dehydrogenation of the methyl group to carboxylate.

F

F

H F

HO HO

H

OH

F

F

H OH H OH

OH

FIGURE 9.28 Transformation of 1-fluoronaphthalene by Cunninghamella elegans. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.)

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Although fluorotoluenes have yielded the corresponding fluorobenzoates, these reactions are merely biotransformations since the fluorobenzoates are terminal metabolites: a. The Pseudomonas sp. strain T oxidized 2-, 3-, and 4-fluorotoluenes to the fluorobenzoates without loss of fluorine, although the yields from 2- and 3-fluorotoluene were low (Seyfried et al. 1994). b. The toluene-degrading organism Azoarcus tolulyticus strain Tol-4 transforms toluene by the fumarate condensation pathway. In the presence of 2-, 3-, or 4-fluorotoluene, the corresponding fluorobenzoates were produced together with the nonfluorinated (E-)phenylitaconate (Chee-Sanford et al. 1996). Presumably, the fluorinated analogs were too transient for unambiguous detection. c. The Thauera sp. strain K172 was capable of dehydrogenating a range of fluorinated toluenes to the corresponding benzoates, including 3- and 4-fluorotoluene and, less effectively, 2-fluorotoluene (Biegert and Fuchs 1995).

FLUOROBENZOATES Aerobic Conditions The metabolism of fluorobenzoates has been examined over many years. Early studies using Nocardia erythropolis (Cain et al. 1968) and Pseudomonas fluorescens (Hughes 1965) showed that although the rates of whole-cell oxidation of fluorobenzoates were less than for benzoate, they were comparable to, and greater than for, the chlorinated analogs. As for their chlorinated analogs, both dioxygenation and hydrolytic pathways may be involved, and studies have revealed that the different pathways depended on the positions of the fluorine substituents. a. The rates of dioxygenation of the fluorobenzoates and chlorobenzoates by Pseudomonas sp. B13 grown with 3-chlorobenzoate were greater for the 3-halogenated compounds than those for the other isomers—though less than that for benzoate (Reineke and Knackmuss 1978). b. Whereas 3-fluoro- and 3-chloro-cis,cis-muconic acids prepared from Pseudomonas sp. B13 by catechol oxidation are cycloisomerized abiotically at pH values between 4 and 5, the corresponding 2-halogenated muconic acids are stable at pH 6 (Schmidt et al. 1980). c. Loss of halogen may take place before or subsequent to ring fission: • For the ortho-fluorinated compounds, dioxygenation may occur with simultaneous decarboxylation and loss of fluoride. • For the 4-fluorinated compounds, loss of fluoride may take place hydrolytically before ring fission. • The pathways for the degradation of 3-halogenated compounds are determined by the inhibitory effect of 3-halogenated catechols, which is discussed in greater detail in Part 1 of this chapter and Chapter 3, Part 1. The metabolism of 2-fluorobenzoate has been described (Goldman et al. 1967) in a strain of Pseudomonas sp. that was able to use this as the sole source of carbon. 3-Fluorocatechol and 2-fluoromuconic acid were identified as metabolites, while studies with 18O2 (Milne et al. 1968) established that both atoms of catechol originated from oxygen. Further details of this reaction were provided by a study of the metabolism of Burkholderia (Pseudomonas) cepacia strain 2CBS, which is able to metabolize 2-fluorobenzoate, 2-chlorobenzoate, 2-bromobenzoate, and 2-iodobenzoate to catechol by concomitant decarboxylation and loss of halide (Figure 9.29) (Fetzner et al. 1992). The inducible 2-halobenzoate 1,2-dioxygenase consisted of two components: an oxygenase (A), which is an iron– sulfur protein with nonidentical subunits, α (Mr 52,000) and β (Mr 20,000) in an α3β3 structure, and a reductase (B), which is an iron–sulfur flavoprotein (Mr 37,500) containing FAD. The enzyme is

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497

O CO 2 H HO

X

C

OH

OH OH

OH

X

FIGURE 9.29 Transformation of 2-halogenated benzoates to catechol by concomitant decarboxylation and dehalogenation. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.) CO 2 H

F

CO 2 H F

F

HO 2 C

CO 2 H CO 2 H

OH NO 2

O O CH(F)CO 2 H CO 2 H

OH F O

CO 2 H CO 2 H

FCH 2 −CO2H +

CH 2−CO2H CH 2−CO2H

FIGURE 9.30 Degradation of 2-fluoro-4-nitrobenzoate by Nocardia erythropolis. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.)

also active for 2,4-difluorobenzoate and less active for 2,6- and 3,4-difluorobenzoate. It is broadly similar to benzoate dioxygenase though different from the FMN-containing dioxygenases. This 2-halobenzoate-1,2-dioxygenase is also different from the three-component enzyme in Pseudomonas aeruginosa strain 142 (Romanov and Hausinger 1994), which contains a [2Fe–2S] ferredoxin, and is active toward 2,5-dichlorobenzoate though less toward 2,6-, 3,4-, and 3,5-dichlorobenzoate. A contrasting pathway was found for the metabolism of 2-fluoro-4-nitrobenzoate by Nocardia erythropolis that does not involve loss of fluorine concomitant with decarboxylation (Cain et al. 1968). The pathway (Figure 9.30) therefore differs from what has subsequently emerged as the principal pathway for metabolism of nonhydroxylated 2-halogenated benzoates. A bacterial strain FLB300 that does not belong to the genus Pseudomonas is able to use 3-fluorobenzoate for growth (Schreiber et al. 1980; Engesser et al. 1990). This strain minimizes the production of toxic 3-fluorocatechol as an intermediate by regioselective distal 1:6 dioxygenation with the formation of 4-fluorocatechol (Figure 9.31). This strategy minimizes the production of the inhibitory 3-fluorocatechol. For maleylacetates that contain a halogen atom (fluorine, chlorine, or bromine) at the 2-position, loss of halogen consumes 1 mol of NADH per mol of substrate, and the further reduction to 3-ketoadipate, an additional mol (Kaschabek and Reineke 1995) (Figure 9.32). A mutant strain of Alcaligenes eutrophus was used (Reiner and Hegeman 1971) to demonstrate that the initial step in the metabolism of 4-fluorobenzoate involves 1:2 dioxygenation to 4-fluoro1,2-dihydrodihydroxybenzene-1-carboxylate. 4-Fluorobenzoate may be used for growth by several bacteria, and metabolites produced by a strain of Pseudomonas sp., which was also able to degrade 4-fluorophenylacetic acid, were identified and used to construct the pathway. Degradation took place by dioxygenation (decarboxylating), intradiol ring fission of the resulting catechol, and loss of fluoride from the muconolactone (Harper and Blakley 1971c). Part of this pathway was confirmed in later studies in which the important role of maleylacetate reductase was determined (Schlömann et al. 1990a,b; Schreiber et al. 1980).

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CO 2 H

CO 2 H 1:2 F

CO 2 H

F F

1:6 HO HO

OH

CO 2 H

CO 2 H CO 2 H

HO

H F

HO 2 CCH 2

O O

F

F F

FIGURE 9.31 Degradation of 3-fluorobenzoate by regioselective dioxygenation. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.) CO 2 H

OH OH

Cl

CO 2 H CO 2 H

Cl

HO 2 C

O O

Cl Cl

HO 2 C

HO 2 C

O O

HO 2 C CH 2 C O

CO 2 H

NADP

CH 2 C O

CO 2 H

3 - Ketoadipate

FIGURE 9.32 Degradation of halogenated maleylacetate to β-ketoadipate in Pseudomonas sp. strain B13. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.)

Strains of Alcaligenes RH021 and RH022 were able to utilize all the monofluorobenzoates for growth, and two degradative pathways for 4-fluorobanzoate were observed: 1. In strain RH025, a hydrolytic mechanism brought about the loss of fluoride to produce 4-hydroxybenzoate, which was further hydroxylated to 3,4-dihydroxybenzoate before intradiol ring fission (Figure 9.33a). 2. In strain RH021, a benzoate dioxygenase produced 4-fluorocatechol, which was degraded by intradiol fission (Figure 9.33b) (Oltmanns et al. 1989). This pathway is analogous to those proposed for the degradation of 4-chlorobenzoate by a strain of Arthrobacter sp. (Marks et al. 1984), for 2,4-dichlorobenzoate by Alcaligenes denitrificans strain NTB-1 after the initial reductive loss of the ortho chlorine substituent (van der Tweel et al. 1987), and in a coryneform bacterium through an intermediate 2,4-dichlorobenzoyl coenzyme A (Romanov and Hausinger 1996). Difluorobenzoates The metabolism of 2,5- and 3,5-difluorobenzoate by Pseudomonas putida strain JT103 has been examined by 19F NMR (Cass et al. 1987). 1:2 Dioxygenation took place with both substrates (Figure 9.34a,b), and for the 2,5-difluorobenzoate this dominated the alternative distal 1:6 dioxygenation

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499

CO 2 H

CO 2 H

CO 2 H

OH

(a)

OH

OH

OH

(b)

OH

F

CO 2 H CO 2 H

F

FIGURE 9.33 Alternative pathways for degradation of 4-fluorobenzoate by Alcaligenes sp. (a) strain RH 025 and (b) strain RH 021. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.) CO 2 H

HO 2 C

OH OH H

(a) F

F

CO 2 H

F

HO 2 C F

F OH

OH OH

OH

F

(b) F

F

F

O

F

O

CO 2 H HO 2 C

HO 2 C

FIGURE 9.34 (a) Transformation of 3,5-difluorobenzoate and (b) degradation of 2,5-difluorobenzoate by Pseudomonas putida JT 102. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.)

and resulted in the production of 4-fluorocatechol, which was then degraded further with loss of the additional fluorine substituent (Figure 9.34b). This pathway was used by Pseudomonas aeruginosa JB2 for the degradation of the analogous 2,5-dichlorobenzoate, which also degraded 2- and 3-chlorobenzoate (Hickey and Focht 1990). Degradation under Denitrifying Conditions The degradation of benzoate under anaerobic denitrifying conditions has been studied in detail and involves the formation of the CoA thioester by a ligase followed by reduction of the ring, hydroxylation, and dehydrogenation before fission. This pathway has been reviewed for both denitrifying and anaerobic phototrophic bacteria (Harwood et al. 1999), although care should be exercised in extrapolating these mechanisms to other groups of organisms. Degradation of 2-fluorobenzoate under denitrifying conditions has been reported without details several times during the past 25 years (Taylor et al. 1979; Schennen et al. 1985; Song et al. 2000a).

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It was shown that by analogy with benzoate itself, benzoyl-coenzyme A synthetase was induced by 2-fluorobenzoate in Pseudomonas sp. strain KB740 under anaerobic conditions and also by 3- and 4-fluorobenzoate that were not, however, degraded (Schennen et al. 1985). A number of bacterial strains isolated by enrichment with halobenzoates have been examined for growth under both aerobic and denitrifying conditions (Song et al. 2000a), and a rather complex pattern of assimilation was observed depending on whether aerobic or anaerobic conditions were employed. This difference was confirmed in a study with a strain designated Thauera aromatica genomovar chlorobenzoica (Song et al. 2000b). The fate of fluorine was not resolved in these studies, and different mechanisms apparently operate under aerobic or denitrifying conditions. Importantly, all the fluorobenzoate isomers were recalcitrant under sulfate-reducing, iron-reducing, and methanogenic conditions (Vargas et al. 2000). This contrasts with the anaerobic degradation of 3-chlorobenzoate, which has been extensively investigated in the sulfate-reducing Desulfomonile tiedjei that synthesizes ATP by dehalogenation (Louie and Mohn 1999). This produces benzoate by reductive dehalogenation of 3-chlorobenzoate, 3-bromobenzoate, and 3-iodobenzoate—but not 3-fluorobenzoate (DeWeerd and Suflita 1990).

FLUORINATED PHENOLS Aerobic Conditions Bacterial Metabolism The aerobic degradation of phenols is initiated by monooxidation with the production of catechols that undergo ring fission. A number of studies have used various strains of Rhodococcus sp. and Mycobacterium sp. to examine the metabolism of fluorinated phenols, and have illustrated important alternatives: 1. The metabolism of pentafluoro-, pentachloro-, and pentabromophenol by Mycobacterium fortuitum strain CG-2 is initiated by a monooxygenase that carries out hydroxylation at the para position (Uotila et al. 1992). Cell extracts of Rhodococcus chlorophenolicus (Mycobacterium chlorophenolicum) strain PCP-1 in the presence of a reductant transformed tetrafluoro-, tetrachloro-, and tetrabromohydroquinone to 1,2,4-trihydroxybenzene by reactions that clearly involve both hydrolytic and reductive loss of fluorine (Uotila et al. 1995). 2. Whole cells of Rhodococcus opacus strain 1cp were used to study the metabolism of fluorophenol isomers (Finkelstein et al. 2000), in which both fluorocatechols and fluoropyrogallols were produced (Figure 9.35a). Both 3- and 4-fluorophenol produced 5-fluoropyrogallol, which was transformed into 2-pyrone-4-fluoro-6-carboxylate (Figure 9.35b). 3. A 19F NMR study used several species of Rhodococcus and a range of mono-, di-, and trifluorophenols. Both hydroxylation and hydrolytic defluorination were observed (Figure 9.36) (Bondar et al. 1998). Metabolism by Yeasts and Fungi The metabolism of fluorinated phenols has been examined quite extensively in yeasts: 1. The metabolism of a range of fluorophenols containing up to five fluorine substituents was examined using phenol hydroxylase from Trichosporon cutaneum (Peelen et al. 1995). Fluorocatechols were formed, with loss of fluoride for some substrates. 2. 19F NMR was used to examine the metabolites produced by the fungus Exophiala jeanselmei. Hydroxylation was observed for 4-fluorophenol, 2,4-difluorophenol, and 2,3,4-trifluorophenol, and the muconates produced from purified phenol hydroxylase and catechol 1,2-dioxygenase are given in Table 9.2 (Boersma et al. 1998).

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Substituted Carbocyclic Aromatic Compounds OH

501

OH F

HO

OH F

HO

F

(a) HO OH

OH F (b)

OH OH

HO

O OH

O

OH HO 2 C F

F

F

F

FIGURE 9.35 Transformation of (a) 2-fluorophenol and (b) 3- and 4-fluorophenol by Rhodococcus sp. strain 1 cp. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.) OH HO

F

F

F

HO 2 C

F CO 2 H

OH F

F

OH OH

F CO 2 H

F

HO 2 C

FIGURE 9.36 Transformation of 2,3-difluorophenol by Rhodococcus opacus strain 135. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.)

TABLE 9.2 Fluorinated Muconates Formed from Fluorophenols by Phenol Hydroxylase and Catechol 1,2-Dioxygenase from Exophilia jeanselmei Fluoromuconate Metabolite 2-Fluoromuconate 2,4-Difluoromuconate 2,3,4-Trifluoromuconate Tetrafluoromuconate

Phenolic Substrate(s) 2- and 3-fluoro; 2,3- and 2,6-difluorophenol 2,4- and 3,5-difluoro; 2,3,5-trifluorophenol 2,3,4- and 3,4,5-trifluorophenol Pentafluorophenol

Anaerobic Conditions The metabolism of phenols under anaerobic conditions has been examined under denitrifying, sulfate-reducing, Fe (III)-reducing, and anaerobic nonmethanogenic conditions. It is plausible to suggest a common pathway that has been elucidated for denitrifying bacteria. This comprises (a) activation of phenol by the formation of phenylphosphate, (b) carboxylation at a position para to

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the hydroxyl group (Breinig et al. 2000), and (c) dehydroxylation of the 4-hydroxybenzoyl CoA to benzoate (Breese and Fuchs 1998). Successive carboxylation and dehydroxylation reactions have been demonstrated with 2- and 3-fluorophenol and resulted in the production of 2- and 3-fluorobenzoate (Genther et al. 1990). Under methanogenic conditions, 6-fluoro-3-methylphenol was degraded to 3-fluorobenzoate via 5-fluoro-4-hydroxy-2-methylbenzoate and 3-fluoro-4-hydroxybenzoate with the apparent loss of a methyl group (Figure 9.37) (Londry and Fedorak 1993).

AROMATIC TRIFLUOROMETHYL COMPOUNDS Compounds bearing trifluoromethyl substituents have a wide range of application, and consistent with their chemical stability, the strongly electron-attracting trifluoromethyl group is recalcitrant to defluorination. During the metabolism of aromatic compounds with trifluoromethyl substituents, this group generally remains intact even after ring fission, although the trifluoromethyl group apparently does not inhibit either ring dioxygenation or ring fission. It is the recalcitrance of the ring-fission product that is responsible for the partial degradation of these substrates. The following illustrative examples are given: 1. The cis-dihydrodiols have been produced from a number of trifluoromethylbenzoates: a. 4-Trifluoromethylbenzoate by Pseudomonas putida strain JTIO7 (Figure 9.38a) (DeFrank and Ribbons 1976) b. 2-Trifluoromethylbenzoate by P. aeruginosa strain 142 (Figure 9.38b) (Selifonov et al. 1995) c. 4-Methyl- and 4-iodotrifluorobenzene by P. putida strain UV-4 (Boyd et al. 1993) 2. Both trifluoromethylbenzoates and trifluoromethylphenols produced 2-hydroxy-6-keto7,7,7-trifluorohepta-2,4-dienoate, which is the immediate ring-fission product. The trifluoromethyl group of the original substrates was retained in the terminal metabolites. OH

OH

F

OH

F

F

CH 3

F

CH 3 CO 2 H

CO 2 H

CO 2 H

FIGURE 9.37 Anaerobic transformation of 6-fluoro-3-methylphenol. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.) CO 2 H

CO 2 H OH H

(a)

H OH CF 3

CF 3

CO 2 H

HO 2 C CF 3

(b)

OH OH F

FIGURE 9.38 Dioxygenation of (a) 4-trifluoromethylbenzoate by Pseudomonas putida strain JT 107 and (b) 2-trifluoromethylbenzoate by Pseudomonas aeruginosa strain 142. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.)

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Substituted Carbocyclic Aromatic Compounds CO 2 H

CO 2 H OH

CO 2 H

503

OH

HO

OH

HO 2 C

H

(a)

COCF 3

H OH CF 3

CF 3

CF 3

CO 2 H

HO 2 C

OH OH HO

(b)

H CF 3

COCF 3 HO 2 C

CF 3

FIGURE 9.39 Metabolism of (a) 4-trifluoromethylbenzoate by Pseudomonas putida strain JT grown with 4-isopropylbenzoate and (b) 3-trifluoromethylbenzoate by Pseudomonas putida strain mt-2 grown with 3methylbenzoate. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.) OH

OH CF 3

OH

HO COCF 3 HO 2 C

CF 3

FIGURE 9.40 Transformation of 2-trifluoromethylphenol by Bacillus thermoleovorans. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3R, Springer Verlag, 2002, pp. 1–74. With permission.)

a. 4-Trifluoromethylbenzoate was cometabolized by 4-isopropylbenzoate-grown cells of P. putida strain JT (Figure 9.39a) (Engesser et al. 1988a). b. 3-Trifluoromethylbenzoate was cometabolized by 3-methylbenzoate-grown cells of P. putida strain mt-2 (Figure 9.39b) (Engesser et al. 1988b). c. The hydroxylation of 2-trifluoromethylphenol to 3-trifluoromethylcatechol by Bacillus thermoleovorans strain A2 was followed by extradiol ring fission (Figure 9.40) (Reinscheid et al. 1998), while the transformation of 3-trifluoromethylcatechol Engesser et al. 1990) produced the stable terminal metabolite.

REFERENCES Anders H-J, A Kaetzke, P Kämpfer, W Ludwig, G Fuchs (1995) Taxonomic position of aromatic-degrading denitrifying pseudomonad strains K 172 and KB 740 and their description as new members of the genera Thauera, as Thauera aromatica sp. nov., and Azoarcus, as Azoarcus evansii sp. nov., respectively, members of the beta subclass of the Protobacteria. Int J Syst Bacteriol 45: 327–333. Biegert T, G Fuchs (1995) Anaerobic oxidation of toluene (analogues) to benzoate (analogues) by whole cells and by cell extracts of a denitrifying Thauera sp. Arch Microbiol 163: 407–417. Boersma MG, TY Dinarieva, WJ Middelhoven, WJH van Berkel, J Doran, J Vervoort, IMCM Rietjens (1998) 19 F nuclear magnetic resonance as a tool to investigate microbial degradation of fluorophenols to fluorocatechols. Appl Environ Microbiol 64: 1256–1263. Bondar VS, MG Boersma, EL Golovlev, J Vervoort, WJH van Berkel, ZI Finkelstein, IP Solyanikova, LA Golovleva, IMCM Rietjens (1998) 19F NMR study on the biodegradation of fluorophenols by various Rhodococcus species. Biodegradation 9: 475–486. Boyd DR, ND Sharma, MV Hand, MR Groocock, NA Kerley, H Dalton, J Chima, GN Sheldrake (1993) Stereodirecting substituent effects during enzyme-catalysed synthesis of cis-dihydrodiol metabolites of 1,4-disubstituted benzene substrates. J Chem Soc Chem Commun 974–976.

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Breese K, G Fuchs (1998) 4-Hydroxybenzoyl-CoA reductase (dehydroxylating) from the denitrifying bacterium Thauera aromatica—prosthetic groups, electron donor, and genes of the molybdenum-flavin-ironsulfur proteins. Eur J Biochem 251: 916–923. Breinig S, E Schiltz, G Fuchs (2000) Genes involved in anaerobic metabolism of phenol in the bacterium Thauera aromatica. J Bacteriol 182: 5849–5863. Cain RB, EK Tranter, JA Darrah (1968) The utilization of some halogenated aromatic acids by Nocardia. Oxidation and metabolism. Biochem J 106: 211–227. Carvalho MF, MIM Ferreira, IS Moreira, PML Castro, DB Janssen (2006) Degradation of fluorobenzene by Rhizobiales strain F11 via ortho cleavage of 4-fluorocatechol and catechol. Appl Environ Microbiol 72: 7413–7417. Cass AEG, DW Ribbons, JT Rossiter, SR Williams (1987) Biotransformation of aromatic compounds monitoring fluorinated analogues by NMR. FEBS Lett 220: 353–357. Cerniglia CE, DW Miller, SK Yang, JP Freeman (1984) Effects of a fluoro substituent on the fungal metabolism of 1-fluoronaphthalene. Appl Environ Microbiol 48: 294–300. Chee-Sanford JC, JW Frost, MR Fries, Z Zhou, JM Tiedje (1996) Evidence for acetyl coenzyme A and cinnamoyl coenzyme A in the anaerobic toluene mineralization pathway in Azoarcus tolulyticus Tol-4. Appl Environ Microbiol 62: 964–873. DeFrank JJ, DW Ribbons (1976) The p-cymene pathway in Pseudomonas putida PL: isolation of a dihydrodiol accumulated by a mutant. Biochem Biophys Res Commun 70: 1129–1135. DeWeerd KA, JM Suflita (1990) Anaerobic aryl reductive dehalogenation of halobenzoates by cell extracts of “Desulfomonile tiedjei.” Appl Environ Microbiol 56: 2999–3005. Engesser KH, MA Rubio, DW Ribbons (1988a) Bacterial metabolism of side chain fluorinated aromatics: cometabolism of 4-trifluoromethyl (TFM)-benzoate by 4-isopropylbenzoate grown Pseudomonas putida JT strains. Arch Microbiol 149: 198–206. Engesser KH, MA Rubio, H-J Knackmuss (1990) Bacterial metabolism of side-chain-fluorinated aromatics: unproductive meta cleavage of 3-trifluoromethylcatechol. Appl Microbiol Biotechnol 32: 600–608. Engesser KH, RB Cain, HJ Knackmuss (1988b) Bacterial metabolism of side chain fluorinated aromatics: cometabolism of 3-trifluoromethyl (TFM)-benzoate by Pseudomonas putida (arvilla) mt-2 and Rhodococcus rubropertinctus N657. Arch Microbiol 149: 188–197. Fetzner S, R Müller, F Lingens (1992) Purification and some properties of 2-halobenzoate 1,2-dioxygenase, a two-component enzyme system from Pseudomonas cepacia 2CBS. J Bacteriol 174: 279–290. Finkelstein ZI, BP Baskunov, MG Boersma, J Vervoort, EL Golovlev, WJH van Berkel, LAA Gololeva, IMCM Rietjens (2000) Identification of fluoropyrogallols as new intermediates in biotransformation of monofluorophenols in Rhodococcus opacus 1 cp. Appl Environ Microbiol 66: 2148–2153. Genther BRS, GT Townsend, PJ Chapman (1990) Effect of fluorinated analogues of phenol and hydroxybenzoates on the anaerobic transformation of phenol to benzoate. Biodegradation 1: 65–74. Goldman P, GWA Milne, MT Pignataro (1967) Fluorine containing metabolites formed from 2-fluorobenzoic acid by Pseudomonas species. Arch Biochem Biophys 118: 178–184. Green NA, AA Meharg, C Till, J Troke, JK Nicholson (1999) Degradation of 4-fluorobiphenyl by mycorrizal fungi as determined by 19F nuclear magnetic resonance spectroscopy and 14C radiolabelling analysis. Appl Environ Microbiol 65: 4021–4027. Harper DB, ER Blakley (1971c) The metabolism of p-fluorobenzoic acid by a Pseudomonas sp. Can J Microbiol 17: 1015–1023. Harwood CS, G Burchardt, H Herrmann, G Fuchs (1999) Anaerobic metabolism of aromatic compounds via the benzoyl-CoA pathway. FEMS Microbiol Revs 22: 439–458. Hickey WJ, DD Focht (1990) Degradation of mono-, di-, and trihalogenated benzoic acids by Pseudomonas aeruginosa JB2. Appl Environ Microbiol 56: 3842–3850. Hughes DE (1965) The metabolism of halogen-substituted benzoic acids by Pseudomonas fluorescens. Biochem J 96: 181–188. Kaschabek SR, W Reineke (1995) Maleylacetate reductase of Pseudomonas sp. strain B13: specificity of substrate conversion and halide elimination. J Bacteriol 177: 320–325. Londry KL, PM Fedorak (1993) Use of fluorinated compounds to detect aromatic metabolites from m-cresol in a methanogenic consortium: evidence for a demethylation reaction. Appl Environ Microbiol 59: 2229–2238. Louie TM, WW Mohn (1999) Evidence for a chemiosmotic model of dehalorespiration in Desulfomonile tiedjei DCB-1. J Bacteriol 181: 40–46.

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Marks TS, ARW Smith, AV Quirk (1984) Degradation of 4-chlorobenzoic acid by Arthrobacter sp. Appl Environ Microbiol 48: 1020–1025. Milne GWA, P Goldman, JL Holzman (1968) The metabolism of 2-fluorobenzoic acid. J Biol Chem 243: 5374–5376. Neilson AH, A-S Allard (2002) Degradation and transformation of organic fluorine compounds. Hand Environ Chem 3N: 137–202. Oltmanns RH, R Müller, MK Otto, F Lingens (1989) Evidence for a new pathway in the bacterial degradation of 4-fluorobenzoate. Appl Environ Microbiol 55: 2499–2504. Peelen S, IMC Rietjens, MG Boersma, J Vervoort (1995) Conversion of phenol derivatives to hydroxylated products by phenol hydroxylase from Trichosporon cutaneum. Eur J Biochem 227: 284–291. Reineke W, H-J Knackmuss (1978) Chemical structure and biodegradability of halogenated aromatic compounds. Substituent effects on 1,2-dioxygenation of benzoic acid. Biochim Biophys Acta 542: 412–423. Reiner AM, GD Hegeman (1971) Metabolism of benzoic acid by bacteria. Accumulation of (-)-3,5-cyclohexadien-1,2-diol,1-carboxylic acid by a mutant strain of Alcaligenes eutrophus. Biochemistry 10: 2530–2536. Reinscheid UM, H Zuilhog, R Müller, J Vervoort (1998) Biological, thermal and photochemical transformation of 2-trifluoromethylphenol. Biodegradation 9: 487–499. Renganathan V (1989) Possible involvement of toluene-2,3-dioxygenase in defluorination of 3-fluorosubstituted benzenes by toluene-degrading Pseudomonas sp. strain T-12. Appl Environ Microbiol 55: 330–334. Romanov V, RP Hausinger (1994) Pseudomonas aeruginosa 142 uses a three-component ortho-halobenzoate 1,2-dioxygenase for metabolism of 2,4-dichloro- and 2-chlorobenzoate. J Bacteriol 176: 3368–3374. Romanov V, RP Hausinger (1996) NADPH-dependent reductive ortho dehalogenation of 2,4-dichorobenzoic acid in Corynebacterium sepedonicum KZ-4 and coryneform bacterium strain NTB-1 via 2,4-dichlorobenzoyl coenzyme A. J Bacteriol 178: 2656–2661. Schennen U, K Braun, H-J Knackmuss (1985) Anaerobic degradation of 2-fluorobenzoate by benzoatedegrading, denitrifying bacteria. J Bacteriol 161: 321–325. Schlömann M, E Schmidt, H-J Knackmuss (1990a) Different types of dienelactone hydrolase in 4-fluorobenzoate-utilizing bacteria. J Bacteriol 172: 5112–5118. Schlömann M, P Fischer, E Schmidt, H-J Knackmuss (1990b) Enzymatic formation, stability, and spontaneous reactions of 4-fluoromuconolactone, a metabolite of the bacterial degradation of 4-fluorobenzoate. J Bacteriol 172: 5119–5129. Schmidt E, G Remberg, H-J Knackmuss (1980) Chemical structure and biodegradability of halogenated aromatic compounds. Halogenated muconic acids as intermediates. Biochem J 192: 331–337. Schmidt S, P Fortnagel, R-M Wittich (1993) Biodegradation and transformation of 4,4′- and 2,4-dihalodiphenyl ethers by Sphingomonas sp. strain SS33. Appl Environ Microbiol 58: 3931–3933. Schmidt S, R-M Wittich, D Erdmann, H Wilkes, W Francke, P Fortnagel (1992) Biodegradation of diphenyl ether and its monohalogenated derivatives by Sphingomonas sp. strain SS3. Appl Environ Microbiol 58: 2744–2750. Schreiber A, M Hellwig, E Dorn, W Reineke, H-J Kmackmuss (1980) Critical reactions in fluorobenzoic acid degradation by Pseudomonas sp. B13. Appl Environ Microbiol 39: 58–67. Seeger M, B Cámara, B Hofer (2001) Dehalogenation, denitrification, dehydroxylation, and angular attack on substituted biphenyls and related compounds by a biphenyl dioxygenase. J Bacteriol 183: 3548–3555. Selifonov SA, JE Gurst, LP Wackett (1995) Regioselective dioxygenation of ortho-trifluoromethylbenzoate by Pseudomonas aeruginosa 142: evidence for 1,2-dioxygenation as a mechanism in ortho-halobenzoate dehalogenation. Biochem Biophys Res Comm 213: 759–767. Seyfried B, G Glod, R Schocher, A Tschech, J Zeyer (1994) Initial reactions in the anaerobic oxidation of toluene and m-xylene by denitrifying bacteria. Appl Environ Microbiol 60: 4047–4052. Song B, NJ Palleroni, MM Häggblom (2000a) Isolation and characterization of diverse halobenzoatedegrading denitrifying bacteria from soils and sediments. Appl Environ Microbiol 66: 3446–3453. Song B, NJ Palleroni, MM Häggblom (2000b) Description of strain 3CB-1, a genomovar of Thauera aromatica, capable of degrading 3-chlorobenzoate coupled to nitrate reduction. Int J Syst Evol Microbiol 50: 551–558. Taylor BF, WL Hearn, S Pincus (1979) Metabolism of monofluoro- and monochlorobenzoates by a denitrifying bacterium. Arch Microbiol 122: 301–306.

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Uotila JS, VH Kitunen, T Coote, T Saastamoinen, M Salkinoja-Salonen, JHA Apajalahti (1995) Metabolism of halohydroquinones in Rhodococcus chlorophenolicus PCP-1 Biodegradation 6: 119–126. Uotila JS, VH Kitunen, T Saastamoinen, T Coote, MM Hägblom, M Salkinoja-Salonen (1992) Characterization of aromatic dehalogenases of Mycobacterium fortuitum CG-2. J Bacteriol 174: 5669–5675. van der Tweel WJJ, JB Kok, JAM de Bont (1987) Reductive dechlorination of 2,4-dichlorobenzoate to 4-chlorobenzoate and hydrolytic dehalogenation of 4-chloro-, 4-bromo-, and 4-iodobenzoate by Acaligenes denitrificans NTB-1. Appl Environ Microbiol 53: 810–815. Vargas C, B Song, M Camps, MM Häggblom (2000) Anaerobic degradation of fluorinated aromatic compounds. Appl Microbiol Biotechnol 53: 342–347.

PART 4 ARENE SULFONATES Anthropogenic arene sulfonates are structural elements of many industrially important dyes, pigments, and anionic surfactants, although they are very seldom encountered as naturally occurring metabolites. Whereas chlorine is only exceptionally removed from aromatic rings during dioxygenation, this generally takes place with carboxyl and sulfonate groups. The pathway for the degradation of aromatic sulfonates has been elucidated in a detailed study (Cain and Farr 1968), and is illustrated for 4-toluenesulfonate in Figure 9.41. The basic reaction is dioxygenation with concomitant elimination of sulfite, which is illustrated for naphthalene-1-sulfonate (Figure 9.42). The dioxygenase has been purified from a naphthalene-2-sulfonate-degrading pseudomonad (Kuhm et al. 1991). The metabolites may, however, be sufficiently reactive to react with NH4+ in the medium and result in the formation of 5-hydroxyquinoline-2-carboxylate from 5-aminonaphthalene-2-sulfonate (Nörtemann et al. 1993). In contrast, when arene sulfonates serve as sources of sulfur in the absence of sulfate, they are degraded by monooxygenation to the corresponding phenol and sulfite (Kertesz 1999). There are a number of alternative pathways: 1. The degradation of 4-toluenesulfonate has been extensively investigated in Comamonas testosteroni, in which the operon tsaMBCD and tsaR have been characterized (Junker et al. 1997). Two pathways were proposed: (a) successive monooxygenation of the methyl group to a carboxyl group before elimination of sulfite by dioxygenation (Locher et al. 1991) (Figure 9.43) and (b) a putatively reductive pathway with initial loss of sulfite (Cook et al. 1999). 2. Considerable attention has been directed to the degradation of naphthalenesulfonates, and an additional reaction has emerged for the elimination of sulfite. The pathway for the CH 3

CH 3

CH 3

CHO CO 2 H

OH SO 3 H

FIGURE 9.41

OH

OH

Biodegradation of 4-toluenesulfonate.

SO 3 H

HO SO 3 H H

OH OH

OH

OH CO 2 H

FIGURE 9.42

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Biodegradation of naphthalene-1-sulfonate.

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Substituted Carbocyclic Aromatic Compounds

507

degradation of naphthalene 1,6- and 2,6-disulfonates involves the expected elimination of the 1- or 2-sulfonate groups, followed by ring fission with the formation of 5-sulfosalicylate. This is then converted into 2,5-dihydroxybenzoate by what may formally be represented as hydroxylation with elimination of sulfite (Figure 9.44) (Wittich et al. 1988). The 1,2dihydroxynaphthalene dioxygenase from Pseudomonas sp. strain (BN6) that degraded naphthalene sulfonates oxidized both 1,2-dihydroxynaphthalene and 2,3-dihydroxybiphenyl (Kuhm et al. 1991), although the same organism also synthesizes a different 2,3-dihydroxybiphenyl dioxygenase that is not, however, involved in the degradation of naphthalene sulfonates (Heiss et al. 1995). 3. Many azo pigments and dyes contain aromatic rings containing sulfonate substituents. They are degraded by reduction and scission of the resulting products to amines. This is discussed in Part 6 of this chapter. 6-Aminonaphthalene-2-sulfonate is degraded by Pseudaminobacter salicylatoxidans via 5-aminosalicylate with subsequent ring fission (Hintner et al. 2001). The degradation of 2-aminobenzenesulfonate is carried out by dioxygenation with concomitant deamination to catechol 3-sulfonate, from which sulfite is eliminated during extradiol ring fission and hydrolysis (Figure 9.45a) (Junker et al. 1994). 4. Sulfite may not necessarily be eliminated before ring fission. The degradation of 4-aminobenzenesulfonate by a mixed culture of Hydrogenophaga palleroni and Agrobacterium radiobacter produced 4-sulfocatechol from which the sulfonate was eliminated CH 3

CO 2 H

CO 2 H

OH SO 3 H

SO 3 H

FIGURE 9.43

OH

Alternative pathway for the biodegradation of 4-toluenesulfonate.

SO 3 H

OH OH

HO 3 S

HO 3 S

FIGURE 9.44

OH

HO 3 S

CO 2 H

OH

HO

CO 2 H

Biodegradation of naphthalene-1,6-disulfonate.

SO 3 H

SO 3 H

HO 2 C

OH

OH

NH 2 (a)

OH SO 3 H

CO 2 H

SO 3 H CO 2 H CO 2 H

(b)

O

OH NH 2

FIGURE 9.45

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OH

CO 2 H CH 2 CO 2 H

SO 3 H

O

CH 2 CO 2 H

SO 3 H O

Degradation of (a) 2-aminobenzenesulfonate and (b) 4-aminobenzenesulfonate.

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spontaneously as sulfite from the 2-sulfolactone after ring fission (Figure 9.45b) (Hammer et al. 1996). Although anaerobic desulfonation of arene sulfonates, in which the sulfonate is used as a source of sulfur, has been described in a Clostridium sp., the product was not identified (Denger et al. 1996, 1997).

REFERENCES Cain RB, DR Farr (1968) Metabolism of arylsulphonates by micro-organisms. Biochem J 106: 859–877. Cook AM, H Laue, F Junker (1999) Microbial desulfonation. FEMS Microbiol Rev 22: 399–419. Denger K, AM Cook (1997) Assimilation of sulfur from alkyl and aryl sulfonates by Clostridium spp. Arch Microbiol 167: 177–181. Denger K, MA Kertesz, EM Vock, R Schön, A Mägli, AM Cook (1996) Anaerobic desulfonation of 4-tolylsulfonate and 2-(4-sulfophenyl)butyrate by a Clostridium sp. Appl Environ Microbiol 62: 1526–1530. Hammer A, A Stolz, and H-J Knackmuss (1996) Purification and characterization of a novel type of protocatechuate 3,4-dioxygenase with the ability to oxidize 4-sulfocatechol. Arch Microbiol 166: 92–100. Heiss G, A Stolz, AE Kuhm, C Müller, J Klein, J Altenbuchner, H-J Knackmuss (1995) Characterization of a 2,3-dihydroxybiphenyl dioxygenase from the naphthalenesulfonate-degrading bacterium strain BN6. J Bacteriol 177: 5865–5871. Hintner J-P, C Lechner, U Riegert, AE Kuhm, T Storm, T Reemtsma, A Stolz (2001) Direct ring fission of salicylate by a salicylate 1,2-dioxygenase activity from Pseudaminobacter salicylatoxidans. J Bacteriol 183: 6936–6942. Junker F, R Kiewitz, AM Cook (1997) Characterization of the p-toluenesulfonate operon tsaMBCD and tsaR in Comamonas testosteroni T-2. J Bacteriol 179: 919–927. Junker F, T Leisinger, AM Cook (1994) 3-sulfocatechol 2,3-dioxygenase and other dioxygenases (EC 1.13.11.2 and EC 1-14-12) in the degradative pathways of 2-aminobenzenesulfonic, benzenesulphonic and 4-toluenesulfonic acids in Alcaligenes sp. strain O-1. Microbiology (U.K.) 140: 1713–1722. Kertesz MA (1999) Riding the sulfur cycle—metabolism of sulfonates and sulfate esters by Gram-negative bacteria. FEMS Microbiol Revs 24: 135–175. Kuhm AE, A Stolz, K-L Ngai, H-J Knackmuss (1991) Purification and characterization of a 1,2-dihydroxynaphthalene dioxygenase from a bacterium that degrades naphthalenesulfonic acids. J Bacteriol 173: 3795–3802. Locher HH, T Leisinger, AM Cook (1991) 4-toluene sulfonate methyl-monooxygenase from Comamonas testosteroni: purification and some properties of the oxygenase component. J Bacteriol 173: 3741–3748. Nörtemann B, A Glässer, R Machinek, G Remberg, H-J Knackmuss (1993) 5-Hydroxyquinoline-2-carboxylic acid, a dead-end metabolite from the bacterial oxidation of 5-aminonaphthalene-2-sulfonic acid. Appl Environ Microbiol 59: 1898–1903. Wittich RM, HG Rast, H-J Knackmuss (1988) Degradation of naphthalene-2,6- and naphthalene-1,6-disulfonic acid by a Moraxella sp. Appl Environ Microbiol 54: 1842–1847.

PART 5 AROMATIC COMPOUNDS WITH NITRO SUBSTITUENTS These are readily produced by nitration of aromatic compounds and are important explosives. The amines formed by reduction are able to undergo a number of reactions, and have a wide range of application in the production of agrochemicals, dyestuffs, and pharmaceuticals.

NITROARENES Nitrotoluenes including 2,4,6-trinitrotoluene (TNT) are important components of explosives and several nitroarenes including the antibacterial nitrofurans have established mutagenicity (Purohit and Basu 2000). Substantial effort has been directed to the degradation of nitroarenes, and to their reduction to amines. Although nitroarene reductases, noted in Chapter 3, Part 3, are distributed in a range of biota, the products may not necessarily represent intermediates in the degradation

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Substituted Carbocyclic Aromatic Compounds NO 2

CH 2−CH3

H2 N

N

F3 C

NO 2

FIGURE 9.46

CH 2−CH3

509 NO 2

H2 N F3 C

NO 2

CH 2 −CH3 N

H2 N

H N

N

F3 C

N

CH 3

Biodegradation of dinitramine.

pathway. The amines may undergo a number of reactions that may contribute to their removal by transformation: 1. N-acetylation to neutral acetanilides that are terminal metabolites (Noguera and Freedman 1996) 2. Dimerization of partially reduced intermediates to nitroazoxytoluenes (Bayman and Radkar 1997) 3. Association with organic matter in soil to produce nonextractable residues (Thorn and Kennedy 2002). In addition, nitroarenes are reactive compounds, and may undergo cyclization to form benziminazoles when the nitro group is vicinal to an amino or substituted amine group. This has been demonstrated as one of the pathways during the fungal metabolism of dinitramine (Figure 9.46) (Laanio et al. 1973), and in a mixed bacterial culture, which metabolized 2-nitroaniline to 2-methylbenziminazole (Hallas and Alexander 1983). A wide range of mechanisms are involved in the degradation and transformation of aromatic compounds with nitro substituents. These include reduction of the nitro group, and dioxygenation, monooxygenation, and reduction of the aromatic ring. A review devoted to 2,4,6-trinitrotolune is available (Esteve-Núnez et al. 2001). Reduction of nitro groups Reduction of nitro groups to amines may occur under either aerobic or anaerobic conditions: 1. The complete sequence of reduction products was produced from 2,6-dinitrotoluene by Salmonella typhimurium strain TA 98 (Sayama et al. 1992)—2-nitroso-6-nitrotoluene, 2-hydroxylamino-6-nitrotoluene, and 2-amino-6-nitrotoluene. 2. Reduction of nitroarenes has been demonstrated in species of Clostridium and Eubacterium, and was associated with the reduction in the mutagenic activity of 1-nitropyrene, and 1,3- and 1,6-dinitropyrene (Rafii et al. 1991). 3. Anaerobic extracts of Clostridium acetobutylicum reduced 2,4,6-trinitrotoluene to 2,4-dihydroxylamino-6-nitrobenzene, which underwent an enzymatic Bamberger-type rearrangement (noted later) to 2-amino-4-hydroxylamino-5-hydroxy-6-nitrotoluene (Hughes et al. 1998). This is especially remarkable since the enzymatic activity was not dependent on the presence of nitroaromatic compounds as growth substrates, and the reduction was mediated by a hydrogenase that contained only Fe, which is typically associated with hydrogen production in clostridia (Watrous et al. 2003). Analogous reactions were found for the carbon monoxide dehydrogenase from Clostridium thermoaceticum (Huang et al. 2000). 4. Clostridium bifermentans reduced 2,4,6-trinitrotoluene to 2,4,6-triaminotoluene, and a metabolite was formed by reaction of one of the amino groups with methylglyoxal (Lewis et al. 1996).

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5. Reduction of the nitro group of nitrodiphenylamines was accomplished by strains of Desulfovibrio sp., Desulfococcus sp., and Desulfomicrobium sp. during growth with lactate or benzoate. During incubations with 2-aminodiphenylamine, phenazine and 4-aminoacridine were formed (Drzyzga et al. 1996). The mechanism whereby the C10 of 4-aminoacridine is formed from the growth substrate lactate was not resolved, although it might plausibly be suggested to involve carboxylation—analogous to that of aniline—followed by cyclization and reduction of the 9-hydroxyacridine. 6. Under anoxic conditions, TNT can serve as a terminal electron acceptor (Esteve-Núnez et al. 2000), with utilization of the compound as a source of nitrogen. A number of products were formed by oxidation of the methyl group and loss of nitrite to 4-hydroxybenzoate (Esteve-Núnez and Ramos 1998). Partial reduction to hydroxylamines may also be involved in the degradation of nitroarenes: 1. Reduction to hydroxylamines, rearrangement, and dioxygenation: a. Degradation by partial reduction to hydroxylamines has been demonstrated for a number of compounds: 4-nitrobenzoate (Groenewegen et al. 1992; Hughes and Williams 2001), 4-nitrotoluene (Spiess et al. 1998), and in one of the pathways of nitrobenzene (Somerville et al. 1995). The resulting hydroxylamines undergo a Bamberger rearrangement by a combination of mutases and lyases to ortho-aminophenols (Nadeau et al. 2003), which are then degraded by oxygenation and ring fission (Figure 9.47). The 3-hydroxylaminophenol mutase from Ralstonia eutropha JMP134 grown with 3-nitrophenol as nitrogen source has been purified (Schenzle et al. 1997), and is able to catalyze the rearrangement of a number of substituted aromatic hydroxylamines. The formation of both 2- and 4-hydroxyphenol from hydroxylaminobenzene is formally comparable to the classic Bamberger rearrangement. As an alternative, the intermediate ortho-aminophenols may produce 2-aminophenoxazine-3-ones as terminal metabolites (Spiess et al. 1998; Hughes et al. 2002). The phenoxazinone nucleus in the secondary

H

OH N

NH

OH

NH 2 OH

H

NH 2 OH H

OH N

NH

OH H

CO 2 H OH

CO 2 H

OH

CO 2 H

CO 2 H

FIGURE 9.47

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Aerobic transformation of phenylhydroxylamines.

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Substituted Carbocyclic Aromatic Compounds

511

metabolite grixazone B of Streptomyces griseus is produced from 3-amino-4-hydroxybenzoate by analogous reactions—oxidation to the quinone imine followed by reaction with N-acetylcysteine, and further oxidation and dimerization to grixazone with loss of a carboxyl group (Suzuki et al. 2006a). The biosynthesis of the precursor 3-amino4-hydroxybenzoate is accomplished by the condensation of aspartate 4-semialdehyde and dihydroxyacetone phosphate, followed by a sequence of Mn2+-catalyzed reactions involving dihydroxypiperidine and piperidone intermediates (Suzuki et al. 2006b). b. The biodegradation of nitrobenzene by a strain of Pseudomonas pseudoalcaligenes is initiated by reduction to the phenylhydroxylamine followed by rearrangement to 2-aminophenol (Bamberger rearrangement). This is then degraded by a 2-aminophenol 1,6-dioxygenase (Nishino and Spain 1993; Lendenmann and Spain 1996; Takenaka et al. 1997) to 2-aminomuconic semialdehyde, which is further degraded by dehydrogenation and deamination to 2-hydroxymuconate (4-oxalocrotonate) (Figure 9.48) (He and Spain 1997, 1998). Although nitrosobenzene was not detected as an intermediate, it is a substrate for the nitroreductase, whereas hydroxylaminobenzene is not further reduced by the enzyme (Somerville et al. 1995). In Pseudomonas putida strain HS-12, the gene encoding the hydroxylamino mutase occurs on a plasmid pNB2, whereas the other genes are carried by plasmid pNB1 (Park and Kim 2000). This pathway is also used by a strain of Mycobacterium sp. for the degradation of 4-nitrotoluene (Spiess et al. 1998), with the formation of 3-hydroxy-4-aminotoluene as an intermediate before further degradation. Analogous pathways are used for the degradation of 4-nitrochlorobenzene involving reduction, dioxygenation, and deamination of the 2-amino-5-chloromuconate (Wu et al. 2006). 2. Interactions may take place between metabolites at different levels of reduction. This may plausibly account for the dimerization of partially reduced intermediates to nitroazoxytoluenes (Bayman and Radkar 1997), and the identification of azoxy compounds as biotransformation products of 2,4-dinitrotoluene by the fungus Mucrosporium sp. (Figure 9.49) (McCormick et al. 1978). 3. An unusual reaction involving reductive elimination of nitrite has been observed in cultures of a Pseudomonas sp. that can use 2,4,6-trinitrotoluene as a source of nitrogen (Duque et al. 1993). The substrate is transformed by successive reductive loss of nitro groups with the formation of toluene. Although toluene cannot be metabolized by this strain, it can be degraded by a transconjugant containing the TOL plasmid from Pseudomonas putida.

NO 2

NH 2 OH

FIGURE 9.48

NH 2

NH 2 CO 2 H

CO 2 H

CO 2 H

CHO

CO 2 H

CO 2 H

Degradation of nitrobenzene via hydroxylamine.

CH 3

NO 2

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− :O

CH 3 NO 2

FIGURE 9.49

O

N

CH 3

N +

NO 2

NO 2

Metabolism of 2,4-dinitrotoluene by Mucrosporium sp.

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Reduction of the Ring It has been shown that some reductases can mediate the reduction of nitro groups and of the ring to Meisenheimer-type complexes that are analogous to those that have been observed in nitrophenols (which are described later): a. A purified reductase B from Pseudomonas fluorescens strain I-C that was tolerant of high concentrations of glycerol trinitrate carried out reduction of one of the nitro groups of 2,4,6-trinitrotoluene to hydroxylamines. This was followed by nonenzymatic formation of amino-dimethyl-tetranitrobiphenyls whose formation required the participation of a Meisenheimer complex formed by initial NADPH reduction (Pak et al. 2000). b. The family of flavoprotein oxidoreductases includes those for pentaerythritol tetranitrate, glycerol trinitrate, morphinone, N-ethylmaleamide, and “old yellow enzyme,” although they have variable preference for their potential substrates (Williams et al. 2004). Pentaerythritol tetranitrate reductase reduces 2,4,6-trinitritoluene only slowly, although several members of the group were able to reduce the ring to Meisenheimer-type complexes with branching of the pathway after initial reduction by NADPH (Khan et al. 2004). Dioxygenation Dioxygenation may result in the elimination of nitrite in reactions that are analogous to the elimination of sulfite from aromatic sulfonates, or halogen from 2-halobenzoates. As an alternative to dioxygenation, toluene-3- and toluene-4-monooxygenases can transform nitrobenzene to 4- or 3-nitrophenol, with the former dominating (Fishman et al. 2004): 1. The degradation of nitrobenzene by Comamonas sp. strain JS765 (Nishino and Spain 1995; Lessner et al. 2002) and of 2-nitrotoluene by Pseudomonas sp. strain JS42 (Haigler et al. 1994) is mediated by dioxygenation with the formation of catechol and 3-methylcatechol, respectively. The enzyme involved in the degradation of 2-nitrotoluene by this strain has been purified and shown to consist of three components—an [Fe–S] protein ISP 2NT that serves as the terminal oxygenase, a reductase2NT that may be a flavoprotein, and an electron-transfer protein ferredoxin2NT similar to Rieske-type ferredoxins (An et al. 1994). The dioxygenase involved in the degradation of 2,4-dinitrotoluene by Burkholderia (Pseudomonas) sp. strain DNT also consists of three components—a terminal oxygenase, an [Fe–S] ferredoxin, and a reductase—and is broadly similar to naphthalene dioxygenase (Suen et al. 1996). Nonetheless, although the 2-nitrotoluene dioxygenase from Pseudomonas sp. strain JS42, the 2,4-dinitrotoluene dioxygenase from B. cepacia, and naphthalene dioxygenase from Pseudomonas sp. strain 9816-4 have comparable dioxygenase activities toward naphthalene, they have widely different specificities toward the isomeric mononitrotoluenes and 2,4-dinitrotoluene (Parales et al. 1998). In the same way, nitrocatechols can be formed initially from 1,3-dinitrobenzenes, followed by elimination of nitrite in a second step (Dickel and Knackmuss 1991; Spanggord et al. 1991). 2. The degradation of both 2,4- and 2,6-dinitrotoluene is initiated by dioxygenation with the elimination of nitrite and the formation of nitrocatechols. For the 2,4-isomer, loss of an additional nitrite takes place with the formation of 2-hydroxy-5-methylbenzoquinone (Spanggord et al. 1991), whereas for the 2,6-isomer, the catechol undergoes fission and nitrite is lost during the last stage of the sequence (Figure 9.50) (Nishino et al. 2000). The degradation of 2,4-dinitrotoluene by Burkholderia cepacia strain R34 involves dioxygenation with elimination of nitrite to 4-methyl-5-nitrocatechol that underwent monooxygenation to 2-hydroxy-5-methylquinone. The 2,4,5-trihydroxytoluene that is formed from this is degraded by extradiol fission to pyruvate and methylmalonate semialdehyde (Figure 9.51) (Johnson et al. 2002).

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Substituted Carbocyclic Aromatic Compounds

513

CH 3

CH 3

CH 3 NO 2

NO 2

O

(a) O

HO OH

NO 2

OH

CH 3

CH 3

CH 3

O2 N

NO 2

HO

NO 2

NO 2

O

(b) HO 2C−CO

HO

FIGURE 9.50 Degradation of (a) 2,4-dinitrotoluene and (b) 2,6-dinitrotoluene.

CH 3

CH 3

CH 3 NO 2

NO 2

OH

O

NO 2 O

HO

CH 3

HO OH

OH

OH

CH 3 HO

CHO CO 2 H

CH 3−CH

CHO CO 2 H

+ CH 3 COCO 2 H

OH

FIGURE 9.51

Biodegradation of 2,4-dinitrotoluene.

3. The degradation of 2,4,6-trinitrotoluene under aerobic conditions is strictly dependent on the reduction of one or more nitro groups to amino groups before dioxygenation (Fiorella and Spain 1997; Johnson et al. 2001).

Side-Chain Oxidation The degradation of 4-nitrotoluene by Pseudomonas sp. takes place by oxidation to 4-nitrobenzoate via 4-nitrobenzyl alcohol and 4-nitrobenzaldehyde (Haigler and Spain 1993; Rhys-Williams et al. 1993; James and Williams 1998).

Peroxidase Oxidation The degradation of 2,4-dinitrotoluene by Phanerochaete chrysosporium involves a mechanism completely different from those already outlined, and involves the function of both the manganese-dependent and the lignin peroxidase systems (Valli et al. 1992). The pathway is shown in Figure 9.52, and is reminiscent of that used by the same organism for the degradation of 2,4dichlorophenol (Valli and Gold 1991). The second step is catalyzed by the manganese peroxidase system with elimination of methanol, and the subsequent loss of nitrite with partial de-O-methylation is carried out by the lipid peroxidase system.

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Environmental Degradation and Transformation of Organic Chemicals CH 3

CH 3 NO 2

NO 2

O NH 2

O

NO 2

OH OCH 3

OCH 3

OH

O

NO 2

NO 2

OCH 3

OH

OH

OH NO 2

FIGURE 9.52

O

Biodegradation of 2,4-dinitrotoluene by Phanerochaete chrysosporium.

NITROBENZOATES The degradation of nitrobenzoates displays the same diversity of reactions that has been found for nitroarenes, and the pathway likewise depends on the position of the nitro group: 1. The transformation of nitrobenzoates by reduction has been demonstrated even though the reduced products were not on the direct pathway for biodegradation (Cartwright and Cain 1959). The degradation of 2-nitrobenzoate is, however, initiated by reduction: a. The degradation of 2-nitrobenzoate by Pseudomonas fluorescens strain KU-7 is initiated by partial reduction to the hydroxylamine followed by rearrangement to 2-amino-3-hydroxybenzoate (3-hydroxyanthranilate), which was degraded by extradiol fission. Further reactions involved decarboxylation, deamination, and further decarboxylation to acetaldehyde and pyruvate (Figure 9.53a) (Muraki et al. 2003). b. Alternatively, in Arthrobacter protophormiae, reduction to anthranilate took place, followed by the established pathway for its degradation (Chauhan and Jain 2000). 2. The degradation of 3-nitrobenzoate by Pseudomonas sp. strain JS51 and Comamonas sp. strain JS46 is initiated by dioxygenation to 3,4-dihydroxybenzoate with loss of nitrite (Nadeau and Spain 1995) (Figure 9.53b). 3. The degradation of 4-nitrobenzoate proceeds by a different pathway. During the degradation of 4-nitrobenzoate by Comamonas acidovorans, 4-nitroso and 4-hydroxylaminobenzoate were formed successively, and the latter was then metabolized to 3,4-dihydroxybenzoate + with the elimination of NH4 (Groenewegen et al. 1992) (Figure 9.53c). It should be noted that these cells were not adapted to growth with either 4-aminobenzoate or 4-hydroxybenzoate that are alternative plausible intermediates. A comparable pathway is used by strains of Pseudomonas sp. for the degradation of 4-nitrotoluene that is initially oxidized to 4-nitrobenzoate via 4-nitrobenzyl alcohol and 4-nitrobenzaldehyde (Haigler and Spain 1993; Rhys-Williams et al. 1993; James and Williams 1998).

NITROPHENOLS Nitrophenols are phytotoxic, and dinoseb (6-sec-butyl-2,4-dinitrophenol) has been used as a herbicide, while nitrophenols have been detected in rainwater and plausible mechanisms for their abiotic formation have been proposed (Kohler and Heeb 2003; Vione et al. 2005). The pathway for the degradation of phenols with a single nitro group depends on the position of the substituents, while

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Substituted Carbocyclic Aromatic Compounds

515

CO 2 H

CO 2 H

CO 2 H

NO 2

NHOH

NH 2

(a)

OH CH 3 CHO + CH 3CO−CO 2H

CHO CO 2 H CO 2 H

NH 2

CO 2 H

(b) NO 2

OH OH

CO 2 H

CO 2 H

CO 2 H

NO 2

NHOH

OH

(c) OH

FIGURE 9.53

Degradation of (a) 2-nitrobenzoate, (b) 3-nitrobenzoate, and (c) 4-nitrobenzoate.

that of phenols with two or more nitro groups is carried out by entirely different pathways involving reduction of the ring: 1. A strain of Pseudomonas putida was able to use both 2- and 3-nitrophenol as sources of carbon and nitrogen, with elimination of nitrite from the former and ammonium from the latter (Zeyer and Kearney 1984). The degradation of 2-nitrophenol by Pseudomonas putida B2 is carried out by monooxygenation and elimination of nitrite to produce catechol. The oxygenase had a broad substrate specificity (Zeyer and Kocher 1988), and might involve a quinone analogous to the monooxygenase-catalyzed loss of nitrite from 4-nitrophenol (Spain and Gibson 1991). Nitrated aromatic metabolites including 3-nitrotyrosine (van der Vliet et al. 1997) have been used as markers for the physiological production of nitric oxide–derived oxidants, and the nature of the active species has been shown to be the •NO2 radical, which is produced by peroxidase oxidation of NO2− and H2O2 (Brennan et al. 2002). The degradation of 3-nitrotyrosine by Variovorax paradoxus strain JS171 and Burkholderia sp. strain JS165 has been examined (Nishino and Spain 2006). It involves reactions analogous to those for the degradation of tyrosine—deamination and decarboxylation to 4-hydroxy-3-nitrophenylacetate. This is converted to 3,4-dihydrophenylacetate (homoprotocatechuate) with loss of nitrite in a reaction (Figure 9.54) analogous to that for the degradation of 2-nitrophenol, which has already been noted. 2. For the degradation of 4-nitrophenol, two pathways that depend on the organism have been described, and both of them involve monooxygenation to quinones at different stages. This is analogous to that of 4-methyl-5-nitrocatechol, which is produced from 2,4-dinitrotolune and has already been described. a. In a strain of Moraxella sp., flavin-dependent oxygenation by a particulate enzyme produced successively benzoquinone and hydroquinone, which is degraded by dioxygenation to 3-ketoadipate (Spain and Gibson 1991) (Figure 9.55a). This is in contrast to the dioxygenase-mediated reaction in Pseudomonas putida B2 (Zeyer and Kocher 1988).

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Environmental Degradation and Transformation of Organic Chemicals

CH 2 CH(NH 2 )CO 2 H

CH 2 COCO 2 H

NO 2

CH 2 CO 2 H

NO 2

NO 2

OH

CH 2 CO 2 H

OH

OH

OH OH

FIGURE 9.54 Degradation of 3-nitrotyrosine. OH

O

NO 2 (a)

O

(b)

NO 2

OH

OH OH

FIGURE 9.55

OH

O

OH

OH O

CO 2 H CO 2 H

OH

O

OH

CO 2 H CO 2 H

OH

Degradation of 4-nitrophenol.

b. In Arthrobacter sp. (Jain et al. 1994), Bacillus sphaericus JS905 (Kadiyala and Spain 1998) (Figure 9.55b), and in Rhodococcus opacus SAO 101 (Kitagawa et al. 2004), loss of nitrite from the initially formed 4-nitrocatechol produced 1,2,4-trihydroxy-benzene, which was further degraded by dioxygenation. 3. The degradation of 3-nitrophenol is quite different and is initiated in Pseudomonas putida strain B2 by partial reduction to the hydroxylamine (Meulenberg et al. 1996). This is followed by a Bamberger-type rearrangement to 2-aminohydroquinone catalyzed by a mutase before ring fission of the resulting 1,2,4-trihydroxybenzene (Meulenberg et al. 1996; Schenzle et al. 1997). The 3-hydroxylaminophenol mutase from Ralstonia eutropha JMP134 grown with 3-nitrophenol as nitrogen source has been purified (Schenzle et al. 1999). It is able to catalyze the rearrangement of a number of substituted aromatic hydroxylamines and the formation of both 2- and 4-hydroxyphenol from hydroxylaminobenzene, which is formally comparable to the classic Bamberger rearrangement. 4. The degradation of phenols containing several nitro groups has revealed an unusual reaction. Although this may involve the elimination of nitrite, it does not apparently proceed by any of the pathways illustrated above, and terminal metabolites may be formed by reduction of the aromatic ring. For example, 4,6-dinitrohexanoate is produced from 2,4-dinitrophenol (Lenke et al. 1992), and 2,4,6-trinitrocyclohexanone from 2,4,6-trinitrophenol (Lenke and Knackmuss 1992). Depending on the strain, degradation of 2,4-dinitrophenol may produce 4,6-dinitrohexanoate or 3-nitroadipate, which is further degraded with loss of nitrite (Figure 9.56) (Blasco et al. 1999). Details of the mechanism have revealed the role of Meisenheimer-type hydride complexes, which is consistent with the presence of the strongly electron-withdrawing nitro groups that facilitate what are formally hydride reductions involving reduced coenzyme F420 (Ebert et al. 1999). For example, 2,4,6-trinitrophenol is reduced successively by hydride transfers catalyzed by F420, followed by loss of nitrite, further reduction, and ring fission to 4,6-dinitrohexanoate (Hofmann et al. 2004).

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Substituted Carbocyclic Aromatic Compounds

517 CO 2 H CH 2 NO 2

OH

OH

NO 2

NO 2

NO 2

O2 N

O2N

CO 2 H CO 2 H

H

O2 N

FIGURE 9.56

H

H

Degradation of 2,4-dinitrophenol.

In a wider perspective, the role of these complexes is supported by the fact that purified pentaerythritol tetranitrate reductase from Enterobacter cloacae reduced 2,4,6-trinitrotoluene successively to the mono- and dihydride Meisenheimer complexes (French et al. 1998). The formation of 2,4-dinitrophenol from 2,4,6-trinitrophenol by Rhodococcus erythropolis (Lenke and Knackmuss 1992, 1996; Rieger et al. 1999) and Nocardioides sp. strain CB 22-2 (Behrend and Heesche-Wagner 1999) is again consistent with a pathway involving loss of nitrite from a Meisenheimer-like hydride complex. An analogous complex has been identified from a strain of Mycobacterium sp. during the metabolism of 2,4,6-trinitrotoluene (Vorbeck et al. 1994, 1998).

REFERENCES An D, DT Gibson, JC Spain (1994) Oxidative release of nitrite from 2-nitrotoluene by a three component enzyme system from Pseudomonas sp. strain JS42. J Bacteriol 176: 7462–7467. Bayman P, GV Radkar (1997) Transformation and tolerance of TNT (2,4,6-trinitrotoluene) by fungi. Int Biodet Biorem 39: 45–53. Behrend C, K Heesche-Wagner (1999) Formation of hydride-Meisenheimer complexes of picric acid (2,4,6trinitrophenol) and 2,4-dinitrophenol during mineralization of picric acid by Nocardiodes sp. strain CB 22-2. Appl Environ Microbiol 65: 1372–1377. Blasco R, E Moore, V Wray, D Pieper, K Timmis, F Castillo (1999) 3-Nitroadipoate, a metabolic intermediate for mineralization of 2,4-dinitrophenol by a new strain of a Rhodococcus species. J Bacteriol 181: 149–152. Brennan M-L and et al. (2002) A tale of two controversies. Defining both the role of peroxidases in nitrotyrosine formation in vivo using eosinophil peroxidase and myeloperoxidase-deficient mice, and the nature of the peroxidase-generated reactive nitrogen species. J Biol Chem 277: 17415–17427. Cartwright NJ, RB Cain (1959) Bacterial degradation of the nitrobenzoic acids. 2. Reduction of the nitro group. Biochem J 73 : 305–314. Chauhan A, RK Jain (2000) Degradation of o-nitrobenzoate via anthranilic acid (o-aminobenzoate) by Arthroacter protophormiae: a plasmid-encoded new pathway. Biochem Biophys Res Commun 267: 236–244. Dickel OD, H-J Knackmuss (1991) Catabolism of 1,3-dinitrobenzene by Rhodococcus sp. QT-1. Arch Microbiol 157: 76–79. Drzyzga O, A Schmidt, K-H Blotevogel (1996) Cometabolic transformation and cleavage of nitrodiphenylamines by three newly isolated sulfate-reducing bacterial strains. Appl Environ Microbiol 62: 1710–1716. Duque E, A Haidour, F Godoy, JL Ramos (1993) Construction of a Pseudomonas hybrid strain that mineralizes 2,4,6-trinitrotoluene. J Bacteriol 175: 2278–2283. Ebert S, P-G Rieger, H-J Knmackmuss (1999) Function of coenzyme F420 in aerobic catabolism of 2,4,6-trinitrophenol and 2,4-dinitrophenol by Nocardioides simplex FJ2-1A. J Bacteriol 181: 2669–2694. Esteve-Núnez A, A Caballero, JL Ramos (2001) Biological degradation of 2,4,6-trinitrotoluene. Microbiol Mol Biol Revs 65: 335–352. Esteve-Núnez A, G Lucchesi, B Phillip, B Schink, JL Ramos (2000) Respiration of 2,4,6-trinitrotoluene by Pseudomonas sp. strain JLR11. J Bacteriol 182: 1352–1355. Esteve-Núnez A, JL Ramos (1998) Metabolism of 2,4,6-trinitrotoluene by Pseudomonas sp. JRL11. Environ Sci Technol 32: 3802–3808.

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Fiorella PD, JC Spain (1997) Transformation of 2,4,6-trinitrotoluene by Pseudomonas pseudoalcaligenes JS52. Appl Environ Microbiol 63: 2007–2015. Fishman A, Y Tao, TK Wood (2004) Toluene 3-monooxygenase of Ralstonia pickettii PKO1 is a para-hydroxylating enzyme. J Bacteriol 186: 3117–3123. French CE, S Nicklin, NC Bruce (1998) Aerobic degradation of 2,4,6-trinitrotoluene by Enterobacter cloacae PB2, and by pentaerythritol tetranitrate reductase. Appl Environ Microbiol 64: 2864–2868. Groenewegen PEG, P Breeuwer, JMLM van Helvoort, AAM Langenhoff, FP de Vries, JAM de Bont (1992) Novel degradative pathway of 4-nitrobenzoate in Comamonas acidovorans NBA-10. J Gen Microbiol 138: 1599–1605. Haigler BE, JC Spain (1993) Biodegradation of 4-nitrotoluene by Pseudomonas sp. strain 4NT. Appl Environ Microbiol 59: 2239–2243. Haigler BE, WH Wallace, JC Spain (1994) Biodegradation of 2-nitrotoluene by Pseudomonas sp. strain JS 42. Appl Environ Microbiol 60: 3466–3469. Hallas LE, M Alexander (1983) Microbial transformation of nitroaromatic compounds in sewage effluent. Appl Environ Microbiol 45: 1234–1241. He Z, JC Spain (1997) Studies of the catabolic pathway of degradation of nitrobenzene by Pseudomonas pseudoalcaligenes JS45: removal of the amino group from 2-aminomuconic semialdehyde. Appl Environ Microbiol 63: 4839–4843. He Z, JC Spain (1998) A novel 2-aminomuconate deaminase in the nitrobenzene degradation pathways of Pseudomonas pseudoalcaligenes JS 45. J Bacteriol 180: 2502–2506. Hofmann KW, H-J Knackmuss, G Heiss (2004) Nitrite elimination and hydrolytic ring cleavage in 2,4,6-trinitrophenol (picric acid) degradation. Appl Environ Microbiol 70: 2854–2860. Huang S, PA Lindahl, C Wang, GN Bennett, FB Rudolph, JB Hughes (2000) 2,4,6-trinitrotoluene reduction by carbon monoxide dehydrogenase from Clostridium thermoaceticum. Appl Environ Microbiol 66: 1474–1478. Hughes JB, C Wang, K Yesland, A Richardson, R Bhadra, G Bennett, F Rudolph (1998) Bamberger rearrangement during TNT metabolism by Clostridium acetobutylicum. Environ Sci Technol 32: 494–500. Hughes MA, MJ Baggs, J al-Dulayymi, MS Baird, PA Williams (2002) Accumulation of 2-aminophenoxazine-3-one-7-carboxylate during growth of Pseudomonas putida TW3 on 4-nitro-substituted substrates requires 4-hydroxylaminbenzoate lyase (PnbB). Appl Environ Microbiol 68: 4965–4970. Hughes MA, PA Williams (2001) Cloning and characterization of the pnb genes, encoding enzymes for 4-nitrobenzoate catabolism in Pseudomonas putida TW3. J Bacteriol 183: 1225–1232. Jain RK, JH Dreisbach, JC Spain (1994) Biodegradation of p-nitrophenol via 1,2,4-benzenetriol by an Arthrobacter sp. Appl Environ Microbiol 60: 3030–3032. James KD, PA Williams (1998) ntn genes determining the early steps in the divergent catabolism of 4-nitrotoluene and toluene in Pseudomonas sp. strain TW3. J Bacteriol 180: 2043–2049. Johnson GR, RK Jain, JC Spain (2002) Origins of the 2,4-dinitrotolune pathway. J Bacteriol 184: 4219–4231. Johnson GR, BF Smets, JC Spain (2001) Oxidative transformation of aminodinitrotoluene isomers by multicomponent dioxygenases. Appl Environ Microbiol 67: 5460–5466. Kadiyala V, JC Spain (1998) A two-component monoxygenase catalyzes both the hydroxylation of p-nitrophenol and the oxidative release of nitrite from 4-nitrocatechol in Bacillus sphaericus JS905. Appl Environ Microbiol 64: 2479–2484. Khan H, T Barna, RJ Harris, NC Bruce, I Barsukov, AW Munro, PCE Moody, NS Scrutton (2004) Atomic resolution structures and solution behavior of enzyme-substrate complexes of Enterobacter cloacae PB2 pentaerythritol tetranitrate reductase. J Biol Chem 279: 30563–30572. Kitagawa W, N Kimura, Y Kamagata (2004) A novel p-nitrophenol degradation gene cluster from a Grampositive bacterium Rhodococcus opacus SAO101. J Bacteriol 186: 4894–4902. Kohler M, NV Heeb (2003) Determination of nitrated phenolic compounds in rain by liquid chromatography/ atmospheric pressure chemical ionization mass spectrometry. Anal Chem 75: 3115–3121. Laanio TL, PC Kearney, DD Kaufman (1973) Microbial metabolism of dinitramine. Pest Biochem Physiol 3: 271-277 Lendenmann U, JC Spain (1996) 2-aminophenol 1,6-dioxygenase: a novel aromatic ring cleavage enzyme purified from Pseudomonas pseudoalcaligenes JS 45. J Bacteriol 178: 6227–6232. Lenke H, H-J Knackmuss (1992) Initial hydrogenation during catabolism of picric acid by Rhodococcus erythropolis HL 24-2. Appl Environ Microbiol 58: 2933–2937. Lenke H, H-J Knackmuss (1996) Initial hydrogenation and extensive reduction of substituted 2,4-dinitrophenols. Appl Environ Microbiol 62: 784–790.

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Lenke H, DH Pieper, C Bruhn, H-J Knackmuss (1992) Degradation of 2,4-dinitrophenol by two Rhodococcus erythropolis strains, HL 24-1 and HL 24-2. Appl Environ Microbiol 58: 2928–2932. Lessner DJ, GR Johnson, RE Parales, JC Spain, DT Gibson (2002) Molecular characterization and substrate specificity of nitrobenzene dioxygenase from Comamonas sp. strain JS 765. Appl Environ Microbiol 68: 634–541. Lewis TA, S Goszczynski, RL Crawford, RA Korus, W Adamassu (1996) Products of anaerobic 2,4,6-trinitrotoluene (TNT) transformation by Clostridium bifermentans. Appl Environ Microbiol 62: 4669–4674. McCormick NG, JH Cornell, AM Kaplan (1978) Identification of biotransformation products from 2,4-dinitrotoluene. Appl Environ Microbiol 35: 945–948. Meulenberg R, M Pepi, JAM de Bont (1996) Degradation of 3-nitrophenol by Pseudomonas putida B2 occurs via 1,2,4-benzenetriol. Biodegradation 7: 303–311. Muraki T, M Taki, Y Hasegawa, H Iwaki, PCK Lau (2003) Prokaryotic homologues of the eukaryotic 3-hydroxyanthranilate 3,4-dioxygenase and 2-amino-3-carboxymuconate-6-semialdehyde decarboxylase in the 2-nitrobenzoate degradation pathway of Pseudomonas fluorescens strain KU-7. Appl Environ Microbiol 69: 1564–1572. Nadeau LJ, Z He, JC Spain (2003) Bacterial conversion of hydroxylamino aromatic compounds by both lyase and mutase enzymes involves intramolecular transfer of hydroxyl groups. Appl Environ Microbiol 69: 2786–2793. Nadeau IJ, JC Spain (1995) The bacterial degradation of m-nitrobenzoic acid. Appl Environ Microbiol 61: 840–843. Nishino SF, GC Paoli, JC Spain (2000) Aerobic degradation of nitrotoluenes and pathway for bacterial degradation of 2,6-dinitrotoluene. Appl Environ Microbiol 66: 2139–2147. Nishino SF, JC Spain (1993) Degradation of nitrobenzene by a Pseudomonas pseudoalcaligenes. Appl Environ Microbiol 59: 2520–2525. Nishino SF, JC Spain (1995) Oxidative pathways for the degradation of nitrobenzene by Comamonas sp. strain JS 765. Appl Environ Microbiol 61: 2308–2313. Nishino SF, JC Spain (2006) Biodegradation of 3-nitrotyrosine by Burkholderia sp. strain JS165 and Variovorax paradoxus JS171. Appl Environ Microbiol 72: 1040–1044. Noguera DR, DL Freedman (1996) Reduction and acetylation of 2,4-dinitrotoluene by a Pseudomonas aeruginosa strain. Appl Environ Microbiol 62: 2257–2263. Pak JW, Kl Knoke, DR Noguera, BG Fox, GH Chambliss (2000) Transformation of 2,4,6-trinitrotoluene by purified xenobiotic reductase B from Pseudomonas fluorescens I-C. Appl Environ Microbiol 66: 4742–4750. Parales JV, RE Parales, SM Resnick, DT Gibson (1998) Enzyme specificity of 2-nitrotoluene 2,3-dioxygenase from Pseudomonas sp. strain JS42 is determined by the C-terminal region of the α subunit of the oxygenase component. J Bacteriol 180: 1194–1199. Park H-S, H-S Kim (2000) Identification and characterization of the nitrobenzene catabolic plasmids pNB1 and pNB2 in Pseudomonas putida. J Bacteriol 182: 573–580. Purohit V, AK Basu (2000) Mutagenicity of nitroaromatic compounds. Chem Res Toxicol 13: 673–692. Rafii F, W Franklin, RH Heflich, CE Cerniglia (1991) Reduction of nitroaromatic compounds by anaerobic bacteria isolated from the human gastrointestinal tract. Appl Environ Microbiol 57: 962–968. Rhys-Williams W, SC Taylor, PA Williams (1993) A novel pathway for the catabolism of 4-nitrotoluene by Pseudomonas. J Gen Microbiol 139: 1967–1972. Rieger P-G, V Sinnwell, A Preuβ, W Francke, H-J Knackmuss (1999) Hydride-Meisenheimer complex formation and protonation as key reactions of 2,4,6-trinitrophenol biodegradation by Rhodococcus erythropolis. J Bacteriol 181: 1189–1195. Sayama M, M Inoue, M-A Mori, Y Maruyama, H Kozuka (1992) Bacterial metabolism of 2,6-dinitrotoluene with Salmonella typhimurium and mutagenicity of the metabolites of 2,6-dinitrotoluene and related compounds. Xenobiotica 22: 633–640. Schenzle A, H Lenke, P Fischer, PA Williams, H-J Knackmuss (1997) Catabolism of 3-nitrophenol by Ralstonia eutropha JMP 134. Appl Environ Microbiol 63: 1421–1427. Schenzle A, H Lenke, JV Spain, H-J Knackmuss (1999) 3-hydroxylaminophenol mutase from Ralstonia eutropha JMP134 catalyzes a Bamberger rearrangement. J Bacteriol 181: 1444–1450. Somerville CC, SF Nishino, JC Spain (1995) Purification and characterization of nitrobenzene nitroreductase from Pseudomonas pseudoalcaligenes JS45. J Bacteriol 177: 3837–3842. Spain JC, DT Gibson (1991) Pathway for degradation of p-nitrophenol in a Moraxella sp. Appl Environ Microbiol 57: 812–819.

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Spanggord RJ, JC Spain, SF Nishino, KE Mortelmans (1991) Biodegradation of 2,4-dinitrotoluene by a Pseudomonas sp. Appl Environ Microbiol 57: 3200–3205. Spiess T, F Desiere, P Fischer, JC Spain, H-J Knackmuss, H Lenke (1998) A new 4-nitrotoluene degradation pathway in a Mycobacterium strain. Appl Environ Microbiol 64: 446–452. Suen W-C, BE Haigler, JC Spain (1996) 2,4-dinitrotoluene dioxygenase from Burkholderia sp strain DNT: similarity to naphthalene dioxygenase. J Bacteriol 178: 4926–4934. Suzuki H, Y Furusho, T Higashi, Y Ohnishi, S Horinouchi (2006a) A novel o-aminophenol oxidase responsible for formation of the phenoxazinone chromophore of grixazone. J Biol Chem 281: 824–833. Suzuki H, Y Ohnishi, Y Furusho, S Sakuda, S Horinouchi (2006b) Novel benzene ring biosynthesis from C3 and C4 primary metabolites by two enzymes. J Biol Chem 281: 36944–36951. Takenaka S, S Murakami, R Shinke, K Hatakeyama, H Yukuwa, K Aoki (1997) Novel genes encoding 2-aminophenol 1,6-dioxygenase from Pseudomonas species AP-3 growing on 2-aminophenol and catalytic properties of the purified enzyme. J Biol Chem 272: 14727–14732. Thorn KA, KR Kennedy (2002) 15N NMR investigation of the covalent binding of reduced TNT amines to soil humic acid, model compounds, and lignocellulose. Environ Sci Technol 36: 3787–3796. Valli K, BJ Brock, DK Joshi, MH Gold (1992) Degradation of 2,4-dinitrotoluene by the lignin-degrading fungus Phanerochaete chrysosporium. Appl Environ Microbiol 58: 221–228. Valli K, MH Gold (1991) Degradation of 2,4-dichlorophenol by the lignin-degrading fungus Phanerochaete chrysosporium. J Bacteriol 173: 345–352. van der Vliet A, JP Eiserich, H Lliwell, BCE Cross (1997) Formation of reactive nitrogen species during peroxidase-catalyzed oxidation of nitrite. J Biol Chem 272: 7617–7625. Vione D, V Maurino, C Minero, E Pelizzetti (2005) Aqueous atmospheric chemistry: formation of 2,4-nitrophenol upon nitration of 2-nitrophenol and 4-nitrophenol in solution. Environ Sci Technol 39: 7921–7931. Vorbeck C, H Lenke, P Fischer, H-J Knackmuss (1994) Identification of a hydride-Meisenheimer complex as a metabolite of 2,4,6-trinitrotoluene by a Mycobacterium strain. J Bacteriol 176: 932–934. Vorbeck C, H Lenke, P Fischer, JC Spain, H-J Knackmuss (1998) Initial reductive reactions in aerobic microbial metabolism of 2,4,6-trinitrotoluene. Appl Environ Microbiol 64: 246–252. Watrous MM, S Clark, R Kutty, S Huang, FB Rudolph, JB Hughes, GN Bennett (2003) 2,4,6-trinitrotoluene reduction by an Fe-only hydrogenase in Clostridium acetobutylicum. J Bacteriol 182: 5683–5691. Williams RE, DA Rathbone, NS Scrutton, NC Bruce (2004) Biotransformation of explosives by the Old Yellow Enzyme family of flavoproteins. Appl Environ Microbiol 70: 3566–3574. Wu J-f, C-y Jiang, B-j Wang, Y-f Ma, Z-p Liu, S-j Liu (2006) Novel partial reductive pathway for 4-chloronitrobenzene and nitrobenzene degradation in Comamonas sp. strain CNB-1. Appl Environ Microbiol 72: 1759–1765. Zeyer J, HP Kocher (1988) Purification and characterization of a bacterial nitrophenol oxygenase which converts ortho-nitrophenol to catechol and nitrite. J Bacteriol 170: 1789–1794. Zeyer J, PC Kearney (1984) Degradation of o-nitrophenol and m-nitrophenol by a Pseudomonas putida. J Agric Food Chem 32: 238–242.

PART 6 AZOARENES Possibly the most significant discovery in the metabolism of aromatic azo compounds had implications that heralded the age of modern chemotherapy. It was shown that the bactericidal effect of the azo dye Prontosil in vivo was in fact due to the action of its transformation product, sulfanilamide, which is an antagonist of 4-aminobenzoate that is required for the synthesis of the vitamin folic acid. Indeed, this reduction is the typical reaction involved in the first stage of the biodegradation of aromatic azo compounds. Aromatic azo compounds, many of which are sulfonated, are components of many commercially important dyes, colorants, and pigments, so that attention has been directed to their degradation and transformation. These compounds are often considered recalcitrant, although their transformation has been accomplished by reduction to amines with scission of the Ar–N=N–Ar bond to produce arylamines. The amines may then be degraded, for example, 6-aminonaphthalene2-sulfonate by dioxygenation and ring fission to 5-aminosalicylate (Haug et al. 1991). This is then

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degraded by dioxygenation and deamination (Stolz and Knackmuss 1993) in a pathway analogous to that for the degradation of salicylate initiated by salicylate-5-hydroxylase (Zhou et al. 2002). Although reduction of the azo groups may be carried out under both aerobic and anaerobic conditions, some of the complicating issues are noted later. Reduction is readily accomplished under anaerobic conditions (Haug et al. 1991), and the azoreductase and nitroreductase from Clostridium perfringens apparently involve the same protein (Rafii and Cerniglia 1993). However, bacterial azoreductases have also been purified from aerobic organisms including strains of Pseudomonas sp. adapted to grow at the expense of azo dyes (Zimmermann et al. 1984), and Staphylococcus aureus (Chen et al. 2005). The amines that result from these reductive transformations then enter well-established metabolic pathways for the degradation of anilines by oxidative deamination to catechol followed by ring fission (McClure and Venables 1986; Fuchs et al. 1991). However, it should be noted that aniline-4-sulfonate (sulfanilate) formed from sulfonated azo dyes may be excreted into the medium, or after further oxygenation, polymerized to terminal products (Kulla et al. 1983). However, sulfanilate is degraded by a mixed culture of Hydrogenophaga palleroni and Agrobacterium radiobacter, and proceeds via catechol-4-sulfonate followed by intradiol ring fission; sulfite is then eliminated from the muconolactone (Hammer et al. 1996). This pathway has been confirmed using an adapted strain of Hydrogenophaga palleroni strain S1 (Blümel et al. 1998). Although azoreductases have been described in many organisms, there are a number of unresolved issues. In Sphingomonas xenophaga BN6, quinoid redox mediators were involved and two of them, 1,2naphthoquinones with 4-amino substituents, were identified during the interrupted degradation of naphthalene-2-sulfonate (Keck et al. 2002). The cells reduced these to hydroquinones that can accomplish reduction of the azo group outside the cells (Rau et al. 2002; Rau and Stolz 2003). The azoreductase from the aerobe Xenophilus azovorans KF46F is a protein lacking metal ions or cofactors, but is able to accept only a limited number of azo compounds of industrial relevance (Blümel et al. 2002), and a thermostable azoreductase from Bacillus sp. strain SF has been described (Maier et al. 2004). A number of factors are, therefore, implicated in the reduction of aromatic azo compounds: a. Limited reduction by whole cells may be the result of limited uptake of the substrates into the cells. b. Flavin-dependent azo reductases may be laboratory artifacts, and do not play a significant role in vivo (Russ et al. 2000). c. Mediators that have already been noted may play an essential role. Reduction of the azo group in dyes with concomitant decolorization—though not necessarily degradation—has been observed in a number of organisms. These include the yeast Issatchenkia occidentalis under microaerophilic conditions in the presence of glucose or ethanol (Ramalho et al. 2004). In this organism, the 1-amino-2-naphthol that was produced could be used as a source of both carbon and nitrogen, whereas N,N′-dimethylaniline served only as a source of nitrogen. Decolorization of sulfonated azo dyes has also been accomplished with fungal peroxidases, and involves oxidative scission of the azo bond followed by a complex set of redox and hydrolytic reactions (Goszczynski et al. 1994). Although laccases have been proposed for the degradation of azo dyes, it has been shown that the enzyme from Trametes villosa produced metabolites that combined with the dyes to produce oligomeric products that retained the azo group, and were therefore unacceptable for discharge (Zille et al. 2005).

REFERENCES Blümel S, M Contzen, M Lutz, A Stolz, H-J Knackmnuss (1998) Isolation of a bacterial strain with the ability to utilize the sulfonated azo compound 4-carboxy-4′-sulfoazobenzene as the sole source of carbon and energy. Appl Environ Microbiol 64: 2315–2317. Blümel S, H-J Knackmuss, A Stolz (2002) Molecular cloning and characterization of the gene coding for the aerobic azoreductase from Xenophilus azovorans KF46. Appl Environ Microbiol 68: 3948–3955.

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Chen H, SL Hopper, CE Cerniglia (2005) Biochemical and molecular characterization of an azoreductase from Staphylococcus aureus, a tetrametric NADPH-dependent flavoprotein. Microbiology (UK) 151: 1433–1441. Fuchs K, A Schreiner, F Lingens (1991) Degradation of 2-methylaniline and chlorinated isomers of 2-methylaniline by Rhodococcus rhodochrous strain CTM. J Gen Microbiol 137: 2033–2039. Goszczynski S, A Paszczynski, MB Pasti-Grigsby, RL Crawford, DL Crawford (1994) New pathways for degradation of sulfonated azo dyes by microbial peroxidases of Phanerochaete chrysosporium and Streptomyces chromofuscus. J Bacteriol 176: 1339–1347. Hammer A, A Stolz, H-J Knackmuss (1996) Purification and characterization of a novel type of protocatechuate 3,4-dioxygenase with the ability to oxidize 4-sulfocatechol. Arch Microbiol 166: 92–100. Haug W, A Schmidt, B Nörtemann, DC Hempel, A Stolz, H-J Knackmuss (1991) Mineralization of the sulfonated azo dye mordant Yellow 3 by a 6-aminonaphthalene-2-sulfonate-degrading bacterial consortium. Appl Environ Microbiol 57: 3144–3149. Keck A, J Rau, T Reemtsma, R Mattes, A Stolz, J Klein (2002) Identification of quinonoid redox mediators that are formed during the degradation of naphthalene-2-sulfonate by Sphingomonas xenophaga BN6. Appl Environ Microbiol 68: 4341–4349. Kulla HG, F Klausener, U Meyer, B Lüdeke, T Leisinger (1983) Interference of aromatic sulfo groups in the microbial degradation of the aza dyes Orange I and Orange II. Arch Microbiol 135: 1–7. Maier J, A Kandelbauer, A Erlacher, A Cavaco-Paulo, GM Gübitz (2004) A new alkali-thermostable azoreductase from Bacillus sp strain SF Appl Environ Microbiol 70: 837–844. McClure NC, WA Venables (1986) Adaptation of Pseudomonas putida mt-2 to growth on aromatic amines. J Gen Microbiol 132: 2209–2218. Rafii F, CE Cerniglia (1993) Comparison of the azoreductase and nitroreductase from Clostridium perfringens. Appl Environ Microbiol 59: 1731–1734. Ramalho PA, MH Cardoso, A Cavaco-Paulo, MT Ramalho (2004) Characterization of azo reduction activity in a novel ascomycete yeast strain. Appl Environ Microbiol 70: 2279–2288. Rau J, A Stolz (2003) Oxygen-insensitive nitroreductases NfsA and NfsB of Escherichia coli function under anaerobic conditions as lawsone-dependent azo reductases. Appl Environ Microbiol 69: 3448–3455. Rau J, H-J Knackmuss, A Stolz (2002) Effects of different quinoid redox mediators on the anaerobic reduction of azo dyes by bacteria. Environ Sci Technol 36: 1497–1504. Russ R, J Rau, A Stolz (2000) The function of cytoplasmic flavin reductases in the reduction of azo dyes by bacteria. Appl Environ Microbiol 66: 1429–1434. Stolz A, H-J Knackmuss (1993) Bacterial metabolism of 5-aminosalicylic acid; enzymatic conversion to malate, pyruvate and ammonia. J Gen Microbiol 139: 1019–1025. Zhou N-Y, J Al-Dulayymi, MS Baird, PA Williams (2002) Salicylate 5-hydroxylase from Ralstonia sp. strain U2: a monooxygenase with close relationships to and shared electron transport proteins with naphthalene dioxygenase. J Bacteriol 184: 1547–1555. Zille A, B Górnacka, A Rehorek, A Cavaco-Paulo (2005) Degradation of azo dyes by Trametes villosa laccase over long periods of oxidative conditions. Appl Environ Microbiol 71: 6711–6718. Zimmermann T, F Gasser, HG Kulla, T Leisinger (1984) Comparison of two bacterial azoreductases acquired during adaptation to growth on azo dyes. Arch Microbiol 138: 37–43.

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Heterocyclic Aromatic Compounds

Although pyridines and quinolines were first produced during the carbonization of coal, they are now available by synthesis in quantities that far exceed those by the former. Phosphorylated ribosides of hydroxylated and aminated pyrimidines and purines make up the basic structure of ribonucleic and deoxyribonucleic acids. The polycyclic oxaarenes are plant metabolites, while thiaarenes are primarily important components of high-sulfur petroleum that must be removed.

PART 1 AZAARENES FIVE-MEMBERED MONOCYCLIC AZA, OXA, AND THIAARENES There are large numbers of naturally occurring representatives, especially of pyrrole that include the important polypyrroles (porphyrins and corrins), and the nitropyrrole antibiotics such as pyrrolomycins and pyrroxamycin. Derivatives of furan have been used as fungicides and N-vinylpyrrolidone is an important monomer for the production of blood plasma extenders and for cosmetic applications. On account of the similarity in the pathways for the aerobic degradation of monocyclic furan, thiophene, and pyrrole, all of them are considered here. Anaerobic degradation of furans is discussed in Part 2 of this chapter. Aerobic Conditions The degradation of furan-2-carboxylate (Trudgill 1969), thiophene-2-carboxylate (Cripps 1973), and pyrrrole-2-carboxylate (Hormann and Andreesen 1991) proceeds by initial ring hydroxylation rather than dioxygenation that is used for the analogs annelated with benzene rings. After ring fission, 2-ketoglutarate is produced from all of these compounds (Figure 10.1), and this then enters central anabolic and catabolic pathways. A strain of Xanthobacter tagetidis is able to grow aerobically at the expense of furan-2-carboxylate, thiophene-2-carboxylate, thiophene-2-acetate, and pyrrole-2-carboxylate (Padden et al. 1997), and Vibrio sp. strain YC1 with some of them (Evans and Venables 1990a). The initial steps in the aerobic degradation of furan 2-carboxylate by Pseudomonas putida F2 involved formation of the coenzyme-A ester followed by hydroxylation with production of 2-ketoglutarate (Kitcher et al. 1972), and analogously in Pseudomonas putida Fu1 (Koenig and Andreesen 1990a). The enzyme has been purified and the involvement of a Mo-containing dehydrogenase was suggested (Koenig and Andreesen 1990a,b). Although formally analogous reactions occur in the degradation of pyrrole-2-carboxylate by Arthrobacter strain Py1 (Hormann and Andreesen 1991), the hydroxylation of pyrrole-2-carboxylate in a species of Rhodococcus is carried out by a two-component monooxygenase (Becker et al. 1997). The degradation of histidine has been explored intensively in a number of organisms including Pseudomonas aeruginosa, P. fluorescens, Aerobacter (Enterobacter) aerogenes, and Bacillus subtilis, and both the enzymology and the regulation of the pathway have been elucidated. Degradation is accomplished by four enzymes. It is initiated by the formation of urocanate and ammonia by the activity of a lyase, urocanate is then converted into 4-iminazolone-5-propionate in an unusual reaction catalyzed by urocanase, followed by reduction and fission of the ring to glutamate and formamide (Kaminskas et al. 1970) (Figure 10.2). The mechanism of urocanase has attracted 523

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(a) CO 2 H

O

HO

CO 2 H

O

CO 2 H

O

O

O CO 2 H

HO 2 C

(b) S

CO 2 H

HO

HO2C−CH2−CH

CO 2 H

S

C

N H

CO 2 H

HO

CO 2 H

S

O

CO 2 H SH

CO 2 H

HO 2 C

O

H

(c)

O

CO 2 H

CHO CO 2 H

N H

HO 2 C

NH 2

CO 2 H

FIGURE 10.1 Aerobic degradation of (a) furan-2-carboxylate, (b) thiophene-2-carboxylate, and (c) pyrrole-2-carboxylate.

CH 2 CH(NH 2)CO 2H N

NH

OH 2

N

HO N

CH 2 CH 2 CO 2 H NH

CH−CO 2H

CH

HO H N

NH

NH

HO H N

CH 2 CH 2 CO 2 H N

CH 2CH 2 CO 2H

O N

CH−CH 2CO 2H

NH

HO 2 C

CH−CH 2CH 2CO 2H NH 2

FIGURE 10.2

Aerobic degradation of histidine.

considerable attention, and it has been proposed that NAD+ plays a role as the prosthetic group (Keul et al. 1979). In contrast, the degradation of iminazolacetate by Pseudomonas sp. strain ATCC 11299B is carried out by oxygenation with a flavoprotein monooxygenase (Maki et al. 1969).

INDOLE AND CARBAZOLE Whereas the degradation of the carboxylates of the monocyclic furan, thiophene, and pyrrole is initiated by hydroxylation, degradation of their benzo analogs is generally carried out by dioxygenation. The degradation of the analogs dibenzofuran and dibenzo-[1,4]-dioxin is discussed in Part 2 of this chapter. Indole and 3-Alkylindoles Several pathways have been demonstrated for the aerobic degradation of tryptophan.

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525

Side-Chain Degradation 1. Degradation that is initiated by tryptophanase has been demonstrated in a number of bacteria (DeMoss and Moser 1969), including Escherichia coli (Morino and Snell 1967), Paenibacillus (Bacillus) alvei (Hoch and DeMoss 1972), Enterobacter aerogenes (Kawasaki et al. 1993), and the tryptophanase operon in Escherichia coli and Proteus vulgaris are similar in structure and regulation (Kamath and Yanofsky 1992). The enzyme carries out an α,β-elimination to produce indole, pyruvate, and NH4+ (Figure 10.3a). The reaction is dependent on pyridoxal that binds to the ε-amino group of lysine and is initiated by enzymatic removal of the proton at the α-position of the resulting Schiff base (Vederas et al. 1978). In addition, the enzyme is capable of carrying out other α,β-eliminations including that of l-cysteine to pyruvate, sulfide, and ammonia; and the synthesis of l-tryptophan from l-serine and indole (Newton et al. 1965). 2. In Pseudomonas fluorescens ATCC 29574, oxidation produced a number of C2-metabolites after initial transamination (Figure 10.3b) (Narumiya et al. 1979). 3. In Pseudomonas pyrrocinia, transformation was initiated by oxidation of the side chain to indole-3-acetate that was further degraded to indole-3-carboxylate before decarboxylation to indole (Figure 10.3c) (Lübbe et al. 1983). Ring Dioxygenation The degradation of indole to anthranilate, putatively by dioxygenation, has been proposed for an undetermined Gram-positive organism (Figure 10.4) (Fujioka and Wada 1968). This pathway has been substantiated by the dioxygenation of l-tryptophan to N-formylkynurenine and hydrolysis to kynurenine. Further degradation may then take place by either of two pathways: (a) formation of anthranilate that is degraded by dioxygenation to catechol or (b) cyclization of N-formylkynurenine to CH2CH(NH2)CO2H (a) + CH3CO−CO2H + NH3 N H

N H

CH2CO−CO2H

CH( OH)CHO

CO−CHO

(b) N H

N H

N H CH2CO2H

(c) N H

N H

FIGURE 10.3 Pathways for degradation of l-tryptophan by (a) tryptophanase, (b) deamination and oxidation, and (c) side-chain oxidation and decarboxylation to indole. H

OH

CO 2 H

CO 2H

NH−CO 2H

NH 2

O N H

FIGURE 10.4

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N H

Biodegradation of indole.

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4-hydroxyquinoline-2-carboxylate (kynurenic acid) that is termed the quinoline pathway and is found in pseudomonads. The subsequent steps in the pathway have been studied in Pseudomonas fluorescens ATCC 11299B (Taniuchi and Hayaishi 1963), and in a strain of Aerococcus sp. (Dagley and Johnson 1963). These reactions involved dioxygenation to 4-hydroxyquinoline-2-carboxylate-7,8-dihydrodiol followed by dehydrogenation, extradiol ring fission, and reduction of the α,β-unsaturated carboxylate (Figure 10.5a). In Burkholderia cepacia strain J2315, an alternative degradation of kynurenine involved monooxygenation to 3-hydroxykynurenine followed by hydrolysis to 3-hydroxyanthranilate, ring fission by 3-hydroxyanthranilate 3,4-dioxygenase, and production of 4-oxalocrotonate, and subsequently, pyruvate and acetaldehyde (Figure 10.5b) (Colabroy and Begley 2005). The degradation of tryptophan in mammalian systems is initiated by the same reactions as far as kynurenine, which is hydroxylated to 3-hydroxykynurenine and 3-hydroxyanthranilate. This is followed by ring fission and nonenzymatic cyclization to pyridine 2,3-dicarboxylate (Malherbe et al. 1994). 3-Hydroxyanthranilate is also an intermediate during the bacterial degradation of 2-nitrobenzoate when it is degraded to acetaldehyde and pyruvate (Muraki et al. 2003). The transformation of both indole-2- and -3-carboxylate by dioxygenases produced isatin, and the dimeric coupling products indigo and indirubin (Eaton and Chapman 1995). Indole can be oxidized by Desulfobacterium indolicum to CO2 using sulfate as electron acceptor (Bak and Widdel 1986). The degradation of indole-3-acetate has been examined in Bradyrhizobium japonicum and was initiated by oxidation to isatin that was hydrolyzed to 2-aminophenylglyoxylate and anthranilic acid (Figure 10.6) (Jensen et al. 1995). CH 2CH(NH 2)CO 2H

O

COCH 2−CH(NH 2)CO2H (a) N H

NH−CHO

N H

O

CO2 H OH

O

CO 2 H

HO HO H

N H

CO2 H

HO

OH H

N H

OH

CH2 CH(NH2 )CO2 H

O

CO2 H

NH2

O

O

CO2 H

(b) N H

NHCHO

NH2

CHO

NH2 CO 2 H

NH2

NH 2 OH

O

O

CO2 H

CO2 H

CO2 H

OH

HO2 C

N H

CO2 H

CO2 H

CH3 COCO2 H + CH 3CHO

CHO

FIGURE 10.5 Pathways for degradation of l-tryptophan (a) via 4-hydroxyquinoline-2-carboxylate and (b) 3-hydroxykynurenine.

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Heterocyclic Aromatic Compounds CH 2 CO 2 H

HO

CH 2 CO 2 H

527 O

O N H

N H

CO−CO 2H

CO 2 H

NH 2

NH 2

O N H

FIGURE 10.6 Pathway for degradation of indole-3-acetate.

Carbazole The transformation of carbazole has been examined, and naphthalene 1,2-dioxygenase activity was induced in Pseudomonas sp. strain NCIB 9816-4 and in Beijerinckia sp. strain B8/36 (Resnick et al. 1993). The 3-hydroxycarbazole that was formed could have resulted from two pathways: (a) from the initial production of cis-3,4-dihydro-3,4-dihydroxycarbazole followed by dehydration or (b) by monooxygenation that cannot be excluded since monooxygenase activity can be mediated by naphthalene 1,2-dioxygenase (Gibson et al. 1995). The degradation of carbazole has been examined in a number of organisms. These include (a) the Gram-negative Pseudomonas sp. strain LD2 (Gieg et al. 1996) and strain CA10 in which the genes occur on a megaplasmid (Nojiri et al. 2001), Sphingomonas sp. strain KA1 (Urata et al. 2006) and (b) the Gram-positive Nocardioides aromaticivorans strain IC177 in which the genes occur in an operon (Inoue et al. 2006). All of them use essentially the same pathway involving dioxygenation, ring fission, and hydrolysis. The degradation by Pseudomonas sp. strain LD2, which uses carbazole as sole source of carbon, nitrogen, and energy, revealed a complex set of intermediates (Gieg et al. 1996). Anthranilic acid and catechol were intermediates, together with a number of terminal transformation products including indole-3-acetate. These products may plausibly arise from 1,2-dioxygenation (Figure 10.7a) or from angular 1,9a-dioxygenation (Figure 10.7b). In both of them production of the dihydrodiol is apparently followed by elimination rather than dehydrogenation, and for angular 1,9-dioxygenation there are several branches in the subsequent pathways (Figure 10.7c,d). Carbazole 1,9a-dioxygenase, which produced biphenyl-2′-amino-2,3-diol, has been examined in Pseudomonas sp. strain CA10, and consisted of a terminal dioxygenase, a ferredoxin, and a ferredoxin reductase (Sato et al. 1997b), which were products of the genes carAa, carAc, and carAd (Nam et al. 2002). The extradiol fission enzyme and the enzyme that produced anthranilic acid by hydrolytic fission of the vinylogous 1:3-diketone were products of carB and carC, respectively (Sato et al. 1997a). The threecomponent dioxygenase (Nam et al. 2002) has a relaxed specificity that includes dioxygenation not only of these heterocyclic compounds but also of the carbocyclic naphthalene, biphenyl, anthracene, and fluoranthene (Nojiri et al. 1999).

PYRIDINE Azaarenes are structural elements of coal and petroleum, and are found in products derived from them by pyrolysis and distillation. They are components of creosote, which has been widely used as a timber preservative. Aerobic Conditions Compared with monocyclic aromatic hydrocarbons and the five-membered azaarenes, the pathways used for the degradation of pyridines are less uniform, and this is consistent with the differences in electronic structure and thereby their chemical reactivity. For pyridines, both hydroxylation and dioxygenation that is typical of aromatic compounds have been observed, although these are often accompanied by reduction of one or more of the double bonds in the pyridine ring. Examples are used to illustrate the metabolic possibilities.

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(b)

OH

N OH H H OH

H H

H OH

O OH

N H

N

HO

(d)

(c)

O

OH

OH

O

OH NH2

N H

CO2H

HO

NH2

O

H

N

CO2H

CHO

NH2

O

N H

O

CO2H

CO2H

NH2

N CHO H CO2H

CO2H

OH

FIGURE 10.7 (a)–(d) Degradation of carbazole. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3J, pp. 1–80, Springer, Heidelberg, 1998. With permission.)

N H

(a)

N

CO2H

N H

CH2CO2H

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Heterocyclic Aromatic Compounds

529

Reduction Reduction may be either the initial reaction in degradation or take place at a later stage after hydroxylation: 1. Pyridine is degraded by two Gram-positive organisms that are postulated to initiate degradation by reduction rather than by hydroxylation (dehydrogenation), which is widely used by azaarenes. Nocardia sp. strain Z1 produced succinic semialdehyde by ring fission of 1,4-dihydropyridine between C2 and C3, whereas Bacillus sp. strain 4 produced glutaric dialdehyde by fission of 2,3-dihydropyridine after hydroxylation, and hydrolysis between N and C-2 (Figure 10.8) (Watson and Cain 1975). These metabolites are then incorporated into normal catabolic cycles. A strain of Azoarcus sp., which can degrade pyridine under both aerobic and denitrifying conditions, was presumed to use the first pathway (Rhee et al. 1997). Analogous pathways are used for the aerobic degradation of 3-methylpyridine by Gordonia nitida LKE31 (Lee et al. 2001) and in the initial step in one of the pathways for N-methylisonicotinate (Wright and Cain 1972), which is formed by photolysis of the herbicide paraquat. 2. The degradation of pyridoxal (vitamin B6) by Pseudomonas sp. strain MA-1 involves 3-hydroxy-2-methylpyridine-5-carboxylate as an intermediate, which is then degraded by reduction of the pyridine ring followed by ring fission (Sundaram and Snell 1969; Nyns et al. 1969) (Figure 10.9). The inducible oxygenase contains FAD and the formation of the product apparently includes a reductive step. Whereas experiments with 18O2 showed that this metabolite contained two atoms of 18O, some 18O incorporation was also observed when H218O was used. This suggests that both dioxygenation and a hydroxylation mechanism might be operating (Sparrow et al. 1969). H

H

H

H OH

N H

N H

CO 2H

OH

CHO

N N H

N

CO 2 H NH 2

O

CHO CO 2 H

FIGURE 10.8 Reductive pathways in degradation of pyridine. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3J, pp. 1–80, Springer, Heidelberg, 1998. With permission.)

CHO

CH 2 NH 2 HO

Me

CH 2 OH

Me

N

Me

CO 2 H

FIGURE 10.9

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N

N

HO

Me

CO 2 H

N

Pyridoxal

Pyridoxanine

HO

CO 2 H CH 2 OH

HO

CH 2 HO 2 C O Me

N

CO 2 H

(b) HO 2 C

CH 2 CH 2 CHO

Degradation of pyridoxal (vitamin B6).

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Environmental Degradation and Transformation of Organic Chemicals

3. The degradation of pyridine-4-carboxylate by Mycobacterium sp. strain INA1 takes place by successive hydroxylations before reduction and ring fission (Kretzer and Andreesen 1991) (Figure 10.10). Hydroxylation In contrast, degradation of 2- and 4-hydroxypyridines—which exists as pyridones—generally, proceeds by successive hydroxylations (dehydrogenations) to 2,3- or 2,6-dihydroxypyridines that are subsequently degraded by ring fission. Hydroxypyridines are also formed from pyridine carboxylates after hydroxylation and decarboxylation, and are then degraded to maleate (Figure 10.11). This pathway is used for the alternative degradation of N-methylisonicotinate (Orpin et al. 1972). Both 2-hydroxy- and 3-hydroxypyridine are hydroxylated to 2,5-dihydroxypyridine by strains of Achromobacter sp. (Houghton and Cain 1972). These metabolites are probably, however, formed by different reactions; whereas 3-hydroxypyridine behaves as a true pyridine, addition of H2O across the C6−N1 bond would produce the 2,5-dihydroxy compound; 2-hydroxypyridine is a cyclic amide and hydroxylation apparently occurs at the diagonal position. The degradation of 4-hydroxypyridine is also initiated by hydroxylation and is followed by dioxygenation before ring fission (Figure 10.12) (Watson et al. 1974).

CO 2 H

CO 2 H

CO 2 H

O

HO

N

FIGURE 10.10

N

O

OH

N

CO 2 H CO 2 H

O

Biodegradation of pyridine-4-carboxylate.

OH

N

CO 2 H

HO

N

CO 2 H

CO 2 H N

HO

N

CO 2 H HO

OH

N

HO

N

OH

HO

N

CO 2 H

CO 2 H CO 2 H

OH

CO 2 H

HO

N

OH

OH N N

CH 3

O N

CO 2 H HO

N

FIGURE 10.11 Interaction of pathways for degradation of pyridine-2-carboxylate, pyridine-3-carboxylate, and nicotine. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Springer, Heidelberg, 1998. With permission.)

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Heterocyclic Aromatic Compounds OH

531 OH

OH OH CO 2 H

N

N

N

CO 2 H

O CO−CO 2H

CO 2 H

CH 2 CHO

NH

FIGURE 10.12

Biodegradation of 4-hydroxypyridine.

HO

HO 2 C

CO 2 H

HO 2 C

+ N

OH

N

OH

CHO

FIGURE 10.13

N

O

H−CO 2H

CONH 2

CHO

Biodegradation of 2,5-dihydroxypyridine.

Hydroxylation is also involved in the degradation of all the pyridine carboxylates and the interrelations of these pathways are shown (Figure 10.11): 1. Pyridine-2-carboxylate is hydroxylated in Arthrobacter picolinophilus to 3,6-dihydroxypyridine-2-carboxylate (as for 2-hydroxypyridine) by a particulate enzyme that introduces oxygen from H2O (Tate and Ensign 1974), has a molecular mass of 130 kDa, and contains Mo (Siegmund et al. 1990). Further degradation involves decarboxylation to 2,5-dihydroxypyridine before ring fission. This reaction has been examined using 18O2 and H218O (Gauthier and Rittenberg 1971), and shown to involve the incorporation of both atoms of oxygen, one each into formate and maleamate (Figure 10.13). 2. Analogously, pyridine-3-carboxylate is hydroxylated to 6-hydroxypyridine 3-carboxylate, and then either to 2,5-dihydroxypyridine in Pseudomonas fluorescens N-9 (Behrman and Stanier 1957a), or in a Bacillus sp. to the 2,6-dihydroxy carboxylate (Ensign and Rittenberg 1964). Hydroxylation to 2,6-dihydroxypyridine carboxylate in a strain of Bacillus sp. (now B. niacinii) was unambiguously shown using H218O to show that both atoms of the oxygen incorporated were derived from water (Hirschberg and Ensign 1971b). Pyridine3-carboxylate and 6-hydroxypyridine-3-carboxylate hydroxylases were shown to contain two molecules of flavin and eight atoms of iron per molecule (Hirschberg and Ensign 1971a). The enzymes (nicotinate dehydrogenase and 6-hydroxynicotinate dehydrogenase) in B. niacini were separated, and each contained FAD, an [Fe–S] center, and probably Mo (Nagel and Andreesen 1990), and the ring fission is accomplished by a 5,6-dioxygenase (Gauthier and Rittenberg 1971). The pathway used for the further metabolism of 6-hydroxypyridine-3-carboxylate by Azorhizobium caulinodans involves reduction (Kitts et al. 1989). 3. Pyridine-4-carboxylate is hydroxylated by Mycobacterium sp. strain INA1 to 2,6dihydroxypyridine-4-carboxylate. Two different hydroxylation enzymes were involved and were apparently Mo-dependent (Kretzer and Andreesen 1991). The formation of 2ketoglutarate can, however, be rationalized equally as β-oxidation to hexahydropyridine2,3,6-trione-4-carboxy-CoA ester followed by hydrolysis. 4. The degradation of 6-methylpyridine-3-carboxylate by an organism selected by enrichment with 6-methylpyridine-3-carboxylate takes place by selective hydroxylation at C-2,

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and the organism is also able to hydroxylate pyridine 3-carboxylate at the same position (Tinschert et al. 1997). 5. Advantage has been taken of the electronic similarity of phthalates to pyridine dicarboxylates: a. Strains capable of degrading o-, m-, or p-phthalate could oxidize the analogous pyridine 2,3-, 2,6-, or 2,5-dicarboxylate (Taylor and Amador 1988). b. The degradation of o-phthalate is initiated by dioxygenation to produce the 4,5-dihydrodiol before dehydrogenation and decarboxylation to 3,4-dihydroxybenzoate. In addition, the dioxygenase is able to accept pyridine-2,3-dicarboxylate and pyridine-3,4-dicarboxylate (quinolinate) (Batie et al. 1987). There are two regions of DNA that encode the enzymes for o-phthalate degradation and the gene encoding quinolinate phosphoribosyl transferase is located between them, and insertional knockout mutants with elevated levels of this enzyme enhanced growth on o-phthalate (Chang and Zylstra 1999). The degradation of nicotine has been examined extensively in Arthrobacter nicotinovorans (oxydans) in which it is mediated by a plasmid (Brandsch et al. 1982; Schenk et al. 1998). In strain P34, the first metabolite was 6-hydroxynicotine, and experiments with 18O2 and H218O showed that the oxygen in the hydroxyl group was derived from H2O (Hochstein and Dalton 1965). Nicotine dehydrogenase has a molecular mass of 120,000 and contains FAD, Mo, Fe, and acid-labile sulfur (Freudenberg et al. 1988). Degradation involves a series of reactions: a. Hydroxylation of the N-methylpyrrolidine ring at the benzylic carbon atom by an FADcontaining oxidase (Dai et al. 1968), followed by fission of the resulting carbinolamine. b. Hydrolysis of the β-ketoamide with the formation of γ-N-methylaminobutyrate and 2,6-dihydroxypyridine (Sachelaru et al. 2005). c. Further hydroxylation to 2,3,6-trihydroxypyridine by an FAD, Mo, and [Fe–S] dehydrogenase (Baitsch et al. 2001). d. Fission of the oxygen-labile 2,3,6-pyridine to maleamate (Figure 10.14) (Holmes et al. 1972; Baitsch et al. 2001). A diazodiphenoquinone is also produced as a chemical artifact (Knackmuss and Beckmann 1973). There are, however, a number of unresolved features of the degradation, including the mechanisms for production of 2,6-dihydroxypyridine and oxidative fission of the pyridine ring. There has been interest in the fate of the products from photolysis of the herbicide paraquat (1,1′-dimethyl-4,4′-bipyridylium ion), in particular, N-methylisonicotinate. Its degradation has been examined in a strain of Achromobacter sp. and resembles one of the pathways for the degradation of pyridine. It is initiated by reduction of the ring, followed by ring fission and decarboxylation, with the production of formate, methylamine, and succinate (Figure 10.15a) (Wright and Cain 1972). An alternative pathway in a Gram-positive organism involved hydroxylation at C-2 and loss of the methyl group, followed by further hydroxylation to 2,6-dihydroxypyridine-4-carboxylate that is degraded to maleate (Figure 10.15b) (Orpin et al. 1972). Dioxygenation Dioxygenation is illustrated by three examples: 1. The degradation of pyridine by a species of Bacillus (Watson and Cain 1975) produces formate and succinate (Figure 10.16). 2. The degradation of 2,5-dihydroxypyridine has been examined using 18O2 and H218O (Gauthier and Rittenberg 1971), and was shown to involve the incorporation of both atoms of oxygen, one each into formate and maleamate. 3. Degradation of the toxin mimosine, which is produced by Leucaena glauca and Mimosa pudica, is initiated by the formation of 3-hydroxypyrid-4-one, plausibly by the activity of

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FIGURE 10.15

FIGURE 10.14

HO

N

+

N

OH

HO 2 C

H

CH 3

CH 3

HO

CO 2

CO 2

OH

CH 3

N

N

N CH 3

N

N

CO 2

CHO

CH 3

N

HO

CHO

N

OH

N

CO 2 H

CO 2 H

CHO

CHO

OH

OH

NHMe

CO 2 H

N

CO 2 H

H

HO 2 C

CH 2

CH 2

CO 2 H

CH CONH 2

CH

HO

CO 2 H

H

OH

O NHMe

Alternative pathways for aerobic degradation of 1-methylpyridinium-4-carboxylate (a) Wright and Cain (1972) and (b) Orpin et al. (1972).

(b)

(a)

CO 2

HO

NHMe

CO 2 H

Me

HO

Me

Aerobic degradation of nicotine.

N

N

N

O

Heterocyclic Aromatic Compounds 533

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534

Environmental Degradation and Transformation of Organic Chemicals H

H CHO N H

N

FIGURE 10.16

N H

CHO

CO 2 H CHO

+ H2N−CHO

Biodegradation of pyridine.

an enzyme analogous to tryptophanase (Borthakur et al. 2003). The 3-hydroxypyrid-4one is then degraded to formate and pyruvate by enzymes encoded by an extradiol dioxygenase (pydA) and a hydrolase (pydB) (Figure 10.17) (Awaya et al. 2005). Halogenated Pyridines Only a few investigations have examined the degradation of halogenated pyridines, some of which are the basis of herbicides: 1. Chlorpyrifos (O,O-diethyl-O-(3,5,6-trichlo-2-pyridyl)phosphorothioate) is hydrolyzed by Enterobacter sp. to 3,5,6-trichloropyridin-2-ol (Singh et al. 2004) whose degradation has been described (Feng et al. 1997). 2. 5-Chloro-2-hydroxypyridine-3-carboxylate is a terminal metabolite in the degradation of 3-chloroquinoline-8-carboxylate, but can be degraded by Mycobacterium sp. strain BA to chlorofumarate by reactions analogous to those described above for pyridine carboxylates (Figure 10.18) (Tibbles et al. 1989a). 3. Oxidation of the trichloromethyl group in 2-chloro-6-(trichloromethyl)-pyridine to the corresponding carboxylic acid (Vannelli and Hooper 1992) by Nitrosomonas europaea occurs at high oxygen concentrations during cooxidation of ammonia or hydrazine. In contrast, at low oxygen concentrations in the presence of hydrazine, reductive dechlorination to 2-chloro-6-dichloromethylpyridine takes place (Figure 10.19). 4. Although Pseudomonas fluorescens strain N-9 can oxidize 5-fluoronicotinic acid, it is unable to grow at its expense. The initial reaction—analogous to that for nicotinic acid—is the formation of 5-fluoro-6-hydroxynicotinic acid (Behrman and Stanier 1957b). Although this was oxidized further, the products were not identifiable by the methods then available. Pyruvate was formed from nicotinic acid and it was suggested by analogy that fluoropyruvate could be produced from 5-fluoronicotinic acid, and thence fluoroacetate that would inhibit effective functioning of the TCA cycle as a source of carbon and energy. In contrast, the degradation of nicotinic acid produces maleic acid (Behrman and Stanier 1957a); and 5-fluorinicotinic acid would be expected to yield fluorofumarate whose degradation by fumarase is known to yield oxalacetate and fluoride (Marletta et al. 1982). The partial anaerobic dechlorination of pentachloropyridine to a trichloropyridine of unestablished orientation by the dehalorespiring Desulfitobacterium hafniense (frappieri) PCP-1 has been demonstrated (Dennie et al. 1998). Anaerobic Conditions As for the aerobic degradation of pyridines, hydroxylation of the heterocyclic ring is a key reaction in the anaerobic degradation of azaarenes by clostridia. Whereas in Clostridium barkeri, the end products are carboxylic acids, CO2, and ammonium, the anaerobic sulfate-reducing Desulfococcus niacinii degraded nicotinate completely to CO2 (Imhoff-Stuckle and Pfennig 1983), although the details of the pathway remain incompletely resolved. The metabolism of nicotinate has been extensively studied in clostridia and the details of the pathway (Figure 10.20) have been delineated in a series of studies (Kung and Tsai 1971; Kung and

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FIGURE 10.19

FIGURE 10.18

FIGURE 10.17

O

N OH

CO 2 H

OH

HO

Cl

H

N

O

N

OH

CO 2 H

CO 2H

CHO

Metabolism of 2-chloro-6-trichloromethylpyridine.

Cl3 C

N

HO

Cl

Cl

Degradation of 3-chloro-6-hydroxypyridine-5-carboxylate.

Cl

Degradation of mimosine.

O

N

OH

H N

CH 2CH(NH 2) − C O 2H

Cl2 CH

HO 2 C

N

O

OH

OH

CO 2H

NH2

N

N

Cl O

Cl

Cl

CO 2 H

CO 2 H

NHCHO

O

CHO

H

Cl

CO 2 H

CO 2 H

H−CO 2H + CH 3COCO 2H

Heterocyclic Aromatic Compounds 535

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536

Environmental Degradation and Transformation of Organic Chemicals CO 2 H N

CO 2 H

HO

N

O

CO 2 H N H

CH 3

CO 2 H

CH 3

CO 2 H

CH 2

CO 2 H

CH 3

CO 2 H

CO 2 H HO 2 C

CHO

CO 2 H HO 2 C

CH 2

CH3−CO2H + CH3−CH2−CO2H

FIGURE 10.20 Anaerobic biodegradation of nicotinate.

Stadtman 1971; Kung et al. 1971). Degradation is initiated by hydroxylation of the ring, and the level of nicotinic acid hydroxylase is substantially increased by the addition of selenite to the medium (Imhoff and Andreesen 1979). Nicotinate hydroxylase from Clostridium barkeri contains molybdenum that is coordinated to selenium, which is essential for hydroxylase activity (Gladyshev et al. 1994). The most remarkable feature of the pathway is the mechanism whereby 2-methylene-glutarate is converted into methylitaconate by a coenzyme B12–mediated reaction (Kung and Stadtman 1971).

QUINOLINE AND ISOQUINOLINE It is appropriate to note that (a) 2- and 4-hydroxyquinolines generally exist as quinolones and (b) the term hydroxylation is not strictly accurate, and the enzymes are preferably described as oxidoreductases. They are formally analogous to the dehydrogenases that are involved in the hydroxylation of purines, and the degradation of quinoline and its methyl homologs (quinaldines) has been extensively investigated. As a generalization, it may be stated that degradation is almost always initiated by hydroxylation at C-2 or (C-4) in the hetero ring, and the resulting hydroxyquinolines (quinolones) are then further oxygenated in the carbocyclic ring. Bacterial Metabolism Quinoline Experiments with H218O using Pseudomonas putida strain 86 (Bauder et al. 1990) showed that the oxygen incorporated into quinol-2-one originates from water. The oxidoreductases have been purified from a number of organisms that degrade quinoline or quinaldine (2-methylquinoline). These include P. putida strain 86 (Bauder et al. 1990), Comamonas testosteroni strain 63 (Schach et al. 1995), Rhodococcus sp. strain B1 (Peschke and Lingens 1991), Arthrobacter sp. strain Rü 61a (De Beyer and Lingens 1993), and Agrobacterium sp. strain 1B, which can degrade quinoline-4carboxylate (Bauer and Lingens 1992). The oxidoreductases have a molecular mass of 300–360 kDa, and contain per molecule, eight atoms of Fe, eight atoms of acid-labile S, two atoms of Mo, and two molecules of FAD. The organic component of the molybdenum cofactor is generally molybdopterin cytosine dinucleotide (Hetterich et al. 1991; Schach et al. 1995). Although fission of the quinolones formed by hydroxylation may sometimes be accomplished by hydrolysis, degradation generally involves hydroxylation of the hetero ring followed by dioxygenation and fission of the carbocyclic ring. The distinction between monooxygenation and dioxygenation in quinoline is, however, less clear than is the case for carbocyclic compounds. An illustrative example is provided by Pseudomonas putida strain 86 in which 8-hydroxy-quinol-2-one is produced from quinol-2-one. The enzyme system consists of a reductase that transfers electrons from NADH and contains FAD and a [2Fe–2S] ferredoxin, and a high-molecular-weight oxygenase consisting of six identical subunits and six Rieske [2Fe–2S] clusters. It was suggested that this complex belonged to class IB oxygenases that include benzoate-1,2-dioxygenase. It is therefore possible that this apparent monooxygenase is, in fact, a dioxygenase introducing oxygen at the 8,8a-positions

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537

followed by an elimination (Rosche et al. 1995). This is reminiscent of naphthalene dioxygenase in which both activities are catalyzed by the same enzyme (Gibson et al. 1995). Hydroxylation Degradation generally involves both hydroxylation and dioxygenation. For example, in organisms that degrade quinoline, initial hydroxylation that takes place at C-2 is followed in some organisms by dioxygenation. For example, Comamonas testosteroni degrades quinoline, first by hydroxylation to quinol-2-one, and subsequently, by dioxygenation to the cis-5,6-dihydrodiol (Schach et al. 1995). Most dioxygenases for carbocyclic aromatic compounds are multicomponent enzymes, but in contrast the 2-hydroxyquinoline-5,6-dioxygenase from C. testosteroni is apparently a singlecomponent enzyme. 4-Hydroxyquinoline-2-carboxylate (kynurenic acid) is an intermediate on one of the pathways for the metabolism of l-tryptophan. Its degradation by Pseudomonas fluorescens ATCC 11299B involves dioxygenation to quinol-4-one-2-carboxylate-7,8-dihydrodiol followed by dehydrogenation, extradiol ring fission, and reduction of the α,β-unsaturated carboxylate, which has already been noted (Figure 10.5a) (Dagley and Johnson 1963; Taniuchi and Hayaishi 1963). The possibly ambiguous position of quinol-2-one 8-monooxygenase has already been noted. The degradation of quinoline by P. fluorescens strain 3 and P. putida strain 86, and by a strain of Rhodococcus sp. has been investigated (Schwarz et al. 1989). The initial reaction was hydroxylation of the heterocyclic ring at C-2 for all the organisms examined, although subsequent reactions were different for the Gram-negative and Gram-positive organisms. In the pseudomonads, 8-hydroxyquinol-2-one was metabolized to 8-hydroxycoumarin by an unestablished reaction involving loss of NH3 (Figure 10.21a). In the rhodococcus, however, hydroxylation initially produced the 6-hydroxy compound and, by further hydroxylation, the 5,6-dihydroxyquinol-2-one followed by ring fission by extradiol dioxygenation to the terminal metabolite (Figure 10.21b). The degradation of quinoline-4-carboxylate by Microbacterium sp. strain H2, Agrobacterium sp. strain 1B, and Pimelobacter simplex strains 4B and 5B follows an analogous pathway to produce 8-hydroxycoumarin-4-carboxylate, which is then hydrolyzed and reduced (Schmidt et al. 1991). The metabolism of 2-methylquinoline in Arthrobacter sp. strain Rü 61a is analogous, with the introduction of oxygen at C-4 (Hund et al. 1990). The enzymes (oxidoreductases) that introduce oxygen into the pyridine rings in Rhodococcus sp. strain B1, Arthrobacter sp. strain Rü61a, and into quinoline in Pseudomonas putida strain 86 are virtually identical. Like those already noted, they have molecular masses of 300–320 kDa and contain Mo, Fe, FAD, and acid-labile sulfur (De Beyer and Lingens 1993). The enzymes from Comamonas testosteroni for hydroxylation of quinoline to 2-hydroxylation (quinoline 2-oxidoreductase), and the dioxygenase responsible for the introduction of oxygen in the benzene ring (2-oxo-1,2-dihydroquinoline 5,6-dioxygenase) have been described (Schach et al. 1995). The hydroxylase showed similarities to the enzyme from the other quinolinedegrading organisms Pseudomonas putida strain 86 and Rhodococcus sp. strain B1. In contrast, the degradation of quinol-4-one by P. putida strain 33/1 involves initial monooxygenation at C-3 (Block and Lingens 1992a), followed by ring fission that is mediated by a dioxygenase with the formation of anthranilate (Block and Lingens 1992b). Dioxygenation For quinoline and its methyl analogs, several types of dioxygenation have been found, and this may take place at the 5,6- or 7,8-positions, followed by fission of the carbocyclic ring (Figure 10.22). For example, a mutant strain of Pseudomonas putida transformed quinoline to the cis-(5R,6S)- and cis-(7S,8R)-dihydrodiols, and quinoxaline to the cis-(5R,6S)-dihydrodiol (Boyd et al. 1992). Dioxygenation at the 7,8-position is involved in the degradation of (a) 3-chloroquinoline8-carboxylate to 5-chloro-2-hydroxynicotinate (Tibbles et al. 1989b) and (b) 4-hydroxyquinoline2-carboxylate (kynurenate), which is an intermediate in one of the degradation pathways of

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N (b) OH

OH

OH

N

N OH

OH

OH H

N H

O

CH 3

HO

HO

N

OH

CH 3

HO

HO

H H

OH

OH

N H

CH 3

OH

CONH2

OH

N

OH

O

OH

O

N

CH 3

CO−CO 2H

HO

OH

HO 2 C

N

OH

O

OH

HO

OH

N

OH

O

CO 2H

O

FIGURE 10.22 Degradation of methylquinolones. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3J, pp. 1–80, Springer, Heidelberg, 1998. With permission.)

H

HO

FIGURE 10.21 Alternative pathways for the degradation of quinoline.

N

(a)

538 Environmental Degradation and Transformation of Organic Chemicals

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Heterocyclic Aromatic Compounds

539

tryptophan. Ring fission of 3-hydroxyquinol-4-one is carried out by an unusual dioxygenation with the incorporation of both atoms of oxygen at C2 and C4, and the formation of anthranilate and carbon monoxide (Fischer et al. 1999). Fungal Metabolism Fluoroquinolones are important antibiotics that have seen extensive use in veterinary medicine, since they are effective against both Gram-negative organisms including Salmonella spp. and Helicobacter (Campylobacter) spp., and Gram-positive organisms. The degradation of the quinolone antibacterial drugs enrofloxacin (Wetzstein et al. 1999) and ciprofloxacin (Wetzstein et al. 1997) has been studied in the brown-rot fungus Gloeophyllum striatum. A wide range of metabolites was isolated, among them a series in which the 6-fluorine substituent was replaced by hydroxyl (Figure 10.23). The major metabolite produced from enrofloxacin by Mucor ramannianus was the N-oxide of the piperazine ring distal from the hetero ring (Parshikov et al. 2000). The important issue of the photolability of aqueous solutions of these antibiotics has been discussed in Chapter 1. Isoquinoline The degradation of isoquinoline by Alcaligenes faecalis strain Pa and Pseudomonas diminuta strain 7 (Röger et al. 1990, 1995) is mediated by an oxidoreductase that produces 1,2-dihydroisoquinoline1-one, followed by ring fission with the formation of o-phthalate and oxidation to 3,4-dihydroxybenzoate (Figure 10.24). The oxidoreductase has been purified and, like most typical azaarene oxidoreductases, contained 0.85 g atoms of Mo per mole, 3.9 g atoms of Fe, and acid-labile S (Lehmann et al. 1994). The oxidoreductase had a rather limited specificity, but it was able to hydroxylate quinazoline (1,3-diazanaphthalene) and phthalazine (2,3-diazanaphthalene) (Stephan et al. 1996). When the position adjacent to the nitrogen atom was substituted with a methyl group, hydroxylation by quinaldine oxidoreductase took place at the 4-position (De Beyer and Lingens 1993). O F

CO 2 H

F

O O

N Et

H2 N

N

N

N

O

N Et

F

CO 2 H

HO

H2 N

N

CO 2 H NH

N

FIGURE 10.23 Transformation of enrofloxacin by Gleophyllum striatum. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3J, pp. 1–80, Springer, Heidelberg, 1998. With permission.) NH 2

CO 2 H CO 2 H

N

NH O

CO 2 H

CO 2 H

OH OH

FIGURE 10.24 Degradation of isoquinoline. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3N, pp. 138–202, Springer, Heidelberg, 2002. With permission.)

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Environmental Degradation and Transformation of Organic Chemicals

PYRIMIDINE AND RELATED COMPOUNDS EXCLUDING PURINES Aerobic Conditions The aerobic degradation of uracil (2,4-dihydroxypyrimidine) takes place by hydroxylation with the production of barbituric acid (2,4,6-trihydroxypyrimidine) or from thymine the analogous 5-methylbarbituric acid. Barbituric acid may then be degraded by hydrolytic reactions to urea and malonate (Hayaishi and Kornberg 1952; Batt and Woods 1961) by a reversal of its chemical synthesis (Figure 10.25). The hydroxylation is formally analogous to that of 2-hydroxypyridine (2-pyridone) to 2,6-dihydroxypyridine, or of quinoline to 2-hydroxyquinoline and 2,8-dihydroxyquinoline. Uracil is formed during the aerobic degradation of 2-amino-6-hydroxypyrimidine (cytidine), and is then hydroxylated to barbituric acid (2,4,6-trihdroxypyrimidine), which is hydrolyzed to urea and malonate (Hayaishi and Kornberg 1952). In contrast, in Streptomyces griseolus cytochrome P450 hydroxylation of the pyrimidine sulfonylureas leaves the ring intact except for hydroxylations (Harder et al. 1991). The degradation of pyrimidines may also take place by reduction of the ring under both aerobic and anaerobic conditions, and is consistent with the fruitful analogy of pyrimidines with nitroarenes. Both the aerobic and anaerobic degradation of the pyrimidines such as uracil, orotic acid, and thymidine take place by reduction to 4,5-dihydro compounds before hydrolytic ring fission to form N-carbamoyl-β-alanine and hydrolysis to aspartate (Xu and West 1992) (Figure 10.26). The dihydropyrimidase that mediates the ring reduction of uracil and thymidine in Pseudomonas stutzeri (Xu and West 1994), and the corresponding reductase for orotase (Ogawa and Shimizu 1995) have been characterized. The degradation or transformation of only a few halogenated azaarenes has been examined under aerobic conditions: 1. 5-Fluoro-2′-deoxyuridine has been extensively used in studies of the mechanism of action of thymidylate synthase, and 5-fluorouracil is an anticancer drug that has provided a lead to the development of others. The metabolism of 5-fluorouracil by the ascomycete fungus Nectria haematococca has been studied using 19F NMR (Parisot et al. 1991). α-Fluoro-β-alanine (2-fluoro-3-aminopropionate) was produced (Figure 10.27), while 5-fluorouridine-5′-mono-, di-, and triphosphate were found in acid extracts of the mycelia, and the 2′- and 3′-monophosphates were recovered from RNA.

O

O

HN O

HN O

N H

H 2 NCONH 2 + CH 2 N

O

CO 2 H CO 2 H

H

FIGURE 10.25 Degradation of uracil.

O

O

HN O

HN N H

CO 2 H

FIGURE 10.26

CRC_7241_ch010.indd 540

O

CO 2 H

H2N N H

CO2 H

O

HO 2C−CH 2−CH(NH2)−CO2H N H

CO 2 H

Reductive degradation of orotic acid.

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Heterocyclic Aromatic Compounds

541

O F

HN O

H2N−CH 2−CH(F)−CO2H N H

FIGURE 10.27 Degradation of 5-fluorouracil by Nectria haematococca.

O

O

(a) O

N H

O

CO 2 H

CO 2 H

H2N

HN

HN

N H

CO 2 H

O

N

CO 2 H

HO 2C−CH 2−CH(NH2)−CO2H + CO 2 + NH 3 O

O

HN

(b) O

HN N H

O

HO 2C−CH 2−CH2−NH−CONH2 N H

FIGURE 10.28 Anaerobic biodegradation of (a) orotic acid and (b) uracil. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3J, pp. 1–80, Springer, Heidelberg, 1998. With permission.)

2. An enzyme may also carry out gratuitous dehalogenation even though this is not its primary function: a. Thymidylate synthetase catalyzes the reductive methylation of 2′-deoxyuridylate to thymidylate using 5,10-methylene-5,6,7,8-tetrahydrofolate as both methyl donor and reductant. The enzyme can also dehalogenate 5-bromo- and 5-iododeoxyuridylate (Wataya and Santi 1975). b. Dihydropyrimidine dehydrogenase is a key enzyme in pyrimidine synthesis, and contains FAD, FMN, and nonheme iron (Lu et al. 1992). It is also able to catalyze the dehalogenation of 5-bromo- and 5-iodo-5,6-dihydrouracil to uracil (Porter 1994). Unusual 2-ketoglutarate-dependent dioxygenations have been found in the yeast Rhodotorula glutinis: a. Conversion of deoxyuridine by 1′-hydroxylation into uracil and ribonolactone (Stubbe 1985). b. Thymine-7-hydroxylase catalyzed the oxidation of the 5-methyl group to 5-hydroxymethyl uracil (Wondrack et al. 1978). Anaerobic Conditions Reduction of the pyrimidine ring has been shown to be the first step in the degradation of orotic acid by Clostridium (Zymobacterium) oroticum (Lieberman and Kornberg 1953, 1954, 1955) (Figure 10.28a), and of uracil by Cl. uracilicum (Campbell 1957a–c, 1960) (Figure 10.28b), which closely follows the pathways for aerobic degradation.

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Environmental Degradation and Transformation of Organic Chemicals

Pyrazine Substituted pyrazines are metabolites of species of the fungus Aspergillus and the bacterium Bacillus. Their degradation has seldom, however, been examined. Degradation by a strain of Mycobacterium sp. was plausibly postulated to proceed by hydroxylation at C-6 followed by ring fission with the elimination of NH4+ (Rappert et al. 2006). Analogous reactions have already been illustrated for the degradation of pyridines. Isoalloxazine Although simple quinoxalines have not been found as natural products, extensive investigation have been devoted to the biodegradation of the important isoalloxazine riboflavin (vitamin B2). Pseudomonas sp. strain RF was obtained by enrichment with riboflavin and a number of degradation products have been characterized. These included oxamide that provided a clue to the pathway and identification of metabolites from labeled substrate (Harkness et al. 1964). The pathway for degradation may plausibly be rationalized on the basis of a number of successive steps: (a) monooxygenation, (b) oxidative loss of the side chain, (c) hydrolysis of the diketoquinoxaline with loss of oxamide, and (d) fission of the benzenoid ring by dioxygenation to 3,4-dimethylpyrone-6-carboxylate (Figure 10.29).

PURINES Aerobic Degradation Whereas the anaerobic degradation of pyrimidines and purines has been extensively examined in a range of organisms, particularly in species of clostridia, aerobic degradation has been studied less often. The aerobic metabolism of uric acid (2,6,8-trihydroxypurine) has been examined in pseudomonads and degradation has been shown to occur by alternative pathways: a. Fission of the pyrimidine ring catalyzed by uricase to produce allantoin and allantoic acid, which is then degraded to glyoxylate and urea (Canellakis and Cohen 1955; Bachrach 1957) (Figure 10.30a). b. Fission of the iminazolone ring to produce barbituric acid, alloxan, and alloxanic acid, which is produced by a benzylic acid rearrangement of the barbituric acid initially produced by monooxygenation (Figure 10.30b) (Hayaishi and Kornberg 1952). The genes encoding the puc genes, which encode the enzymes in the aerobic pathway for the degradation of hypoxanthine, have been determined for Bacillus subtilis (Schultz et al. 2001), and involve the sequences (Figure 10.31): Adenine (6-aminopurine) → hypoxanthine (6-hydroxypurine) → xanthine (2,6-dihydroxypurine) → uric acid (2,6,8-trihydroxypurine) → allantoin → allantoic acid → ureidoglycine → ureidoglycolic acid. The first two reactions are catalyzed by xanthine dehydrogenase, ring fission by urocanase, and subsequent hydrolysis by S(+)allantoinase and allantoate amidohydrolase. Purines are not generally used as nitrogen sources by Escherichia coli and, although allantoate was putatively produced from adenine, neither NH4+ nor carbamoyl phosphate was produced (Xi et al. 2000). This is consistent with the use of allantoin as a nitrogen source under anaerobic conditions by Escherichia coli that is not, however, able to use it as a growth substrate (Cusa et al. 1999). Anaerobic Degradation Demonstration of the anaerobic degradation of purines belongs to the golden age of microbiology and was appropriately discovered in Beijerinck’s laboratory in Delft. Liebert (1909) obtained a pure culture of an organism that was able to grow anaerobically with 2,6,8-trihydroxypurine (uric acid),

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O

FIGURE 10.30

O

N H

Me

N H

O

O

OH

N H

O

O

CO 2 H

O

N H

NH

NH O

CONH2

OH

N H

OH

Me

O

O

Me

O

O C H 2N

O

HN

CHO

(b)

Me

Me

Alternative aerobic degradation of uric acid.

N H

H N

N H

H N

H 2NCONH2 +

(a)

N H

O

HO 2 C H 2N

O

N

H Me

NH

O

Aerobic degradation of riboflavin.

HN

FIGURE 10.29

N H

Me

O

N

N

Me

O

H

R

O

O

HN

CO2 H

NH

O

O

N H

HO H 2N

O

N H O O HO

N

R

N H

O

NH

NH

N

O

O

CONH2

Me

Me

O

H 2N

O

N H

O

OH

OH

O

NH

NH

O

O

HN O

N H

N H

N

R

O

O

O

O

HN N H

O

NH

H

O

O

NH

O

CO 2 H

OH

N H

N

HO 2C H 2N

O

H 2N

Me

Me

R

Heterocyclic Aromatic Compounds 543

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544

Environmental Degradation and Transformation of Organic Chemicals NH2 N N

O

O N N H

N

HN N

N H

HN O

O N

N

N

H

H

N

HN O

H O

N

N

H

H

FIGURE 10.31 Aerobic degradation of hypoxanthine.

which he named Bacillus acidi urici (Leibert 1909). (The organism is now known as Clostridium acidurici.) Subsequently, several other purinolytic clostridia have been isolated, as well as anaerobic bacteria belonging to other groups. Ring-Fission Reactions Substantial effort has been devoted to elucidating the details of the anaerobic degradation of purines containing hydroxyl and amino groups. Although the most studied group of organisms are the nutritionally restricted clostridia, Clostridium acidurici, Cl. cylindrospermum, and Cl. purinilyticum (Schieffer-Ullrich et al. 1984), attention should also be drawn to the nonspore-forming Eubacterium angustum (Beuscher and Andreesen 1984), Peptostreptococcus barnesae (Schiefer-Ullrich and Andreesen 1985), and Methanococcus vannielii, which can use purines as sources of nitrogen (DeMoll and Auffenberg 1993). Barker and his colleagues (Barker 1961) carried out many of the basic investigations on the mechanisms of purine degradation by clostridia, and more recent developments have been presented (Dürre and Andreesen 1983). One of the significant findings was the selenium dependency in Cl. purinilyticum that was the result of its requirement for the synthesis of several critical enzymes— xanthine dehydrogenase, formate dehydrogenase, and glycine reductase. Selenium is required for the synthesis of active xanthine dehydrogenase in Cl. acidiurici and is a component of the purified enzyme that contained iron, acid-labile sulfur, FAD, and molybdenum in the ratios 7.7:7.5:1.7:1.8 (Wagner et al. 1984). Under conditions of selenium starvation in Cl. purinilyticum, uric acid is the central metabolite and is degraded by fission of the iminazole ring to produce 4,5-diaminouracil, which is then degraded to formate, acetate, glycine, and CO2 (Dürre and Andreesen 1982). Under conditions of selenium sufficiency, xanthine is the central metabolite that is produced from all purines including uric acid, which is reductively transformed to xanthine (Dürre and Andreesen 1983). The pathways for the degradation of purines containing amino and/or hydroxyl groups converge on the synthesis of xanthine (2,6-dihydroxypurine), and are followed by its degradation to formiminoglycine (Figure 10.32). This compound is then used for the synthesis of glycine and 5-formiminotetrahydrofolate whose further metabolism to formate results in the synthesis of ATP. The energy requirements of the cell are supplemented by the contribution of ATP produced during the reduction of glycine to acetate in an unusual reaction catalyzed by glycine reductase (Arkowitz and Abeles 1989). An essentially similar pathway is used by Methanococcus vannielii that can utilize a number of purines as nitrogen sources (DeMoll and Auffenberg 1993). There are some important additional issues: 1. The anaerobe Peptococcus (Micrococcus) aerogenes had a dehydrogenase that carried out specific hydroxylation at the 6-positions of 2- and 8-hydroxypurine, and was therefore distinct from xanthine dehydrogenase from which it could be separated (Woolfolk et al. 1970). It was also able to carry out dismutation of 2-hydroxypurine to xanthine (2,6-dihydroxypurine) and hypoxanthine (6-hydroxypurine).

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Heterocyclic Aromatic Compounds

545 OH

N

N

N H

N

N

N

OH N

N

N

N H

N H

N

N

N

OH

OH

HO

OH N H

N

OH N HO

N

N

HO 2 C

N

HO 2 C

N

N H

H2N−CO−N H

N H

H2 N

N H

N

H2 N

N H

N H2N−CH 2

HO 2 C N H

N H

HO

N

CO2H

THFA OH N H2 N

N

CH

NH

N

CH 2

N H

FIGURE 10.32

N NHR

N +

NHR

N H

H N

CHO CH2−N R

HCO 2 H + THFA

N H

Anaerobic biodegradation of purine.

2. Although it had been assumed that only hypoxanthine dehydrogenase is required for the conversion of hypoxanthine (6-hydroxypurine) into uric acid, in Clostridium purinolyticum, two enzymes, both of which contain a selenium cofactor, are required. The enzymes differ in the molecular mass of their subunits, in their terminal amino acid sequences, in their kinetic parameters, and in their specific activities for purines (Self and Stadman 2000). Purine hydroxylase converts purine into hypoxanthine and xanthine (2,6-dihydroxypurine), which is then further hydroxylated to uric acid (2,6,8-trihydroxypurine) by xanthine dehydrogenase (Self 2002).

TRIAZINES 1,3,5-Triazines Derivatives of 1,3,5-triazine are important herbicides so that attention has been directed to their persistence particularly in the terrestrial environment. In some experiments when a growth substrate was supplied, they have been used as sources of nitrogen or sulfur: a. 2-Chloro-4,6-diamino-1,3,5-triazine as a nitrogen source with lactate as carbon source (Grossenbacher et al. 1984) or 2-chloro-4-aminoethyl-6-amino-1,3,5-triazine as a nitrogen source and glycerol as carbon source (Cook and Hütter 1984). b. 2-(1-Methylethyl)amino-4-hydroxy-6-methylthio-1,3,5-triazine as sulfur source and glucose as carbon source (Cook and Hütter 1982). Degradation of substituted triazines is accomplished by a sequence of hydrolytic reactions (Jutzi et al. 1982; Mulbry 1994). Atrazine, which is a representative herbicide, is degraded

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Environmental Degradation and Transformation of Organic Chemicals Cl N

N

RN

OH

OH N

N

H2 N

NHEt

N

N N

HO

NHEt

N N

NH 2 O

OH

NH 2

NH

O

H

FIGURE 10.33 Degradation of atrazine. O

ϕ

N (a)

N

ϕ

NH 2

CO 2 H NH 2 NH N

N

CH 3

N

ϕ

CH 3

CO 2 H NH 2 N N

N H

ϕCO−CO 2H

+ CH 3CO−NH−NH2

CH 3

O (CH 3 )3 C O (CH 3 )3 C (b)

N N

N

NH N

NH 2

O N

SC H 3

(CH 3 )3 C

NH N

SC H 3 O (CH 3 )3 C

N N

N H

N H

O

NH 2 O

FIGURE 10.34 Aerobic degradation of 1,2,4-triazolone herbicides (a) metamitron and (b) metribuzin.

to cyanuric acid (trihydroxytriazine), and ultimately to CO2 and ammonia (Figure 10.33) (Mulbry 1994; Seffernick et al. 2002; Strong et al. 2002; Fruchey et al. 2003). The trihydroxy compound (cyanuric acid) is produced by three hydrolases encoded in Pseudomonas sp. strain ADP by AtzA, AtzB, and AtzC. Biuret, which is the ring fission product of AtzD, is then hydrolyzed by AtzE to allophanate, and by AtzF to NH4+ and CO2 (Cheng et al. 2005; Shapir et al. 2005). 1,2,4-Triazines There are a few important examples of 1,2,4-triazine herbicides. The 1,2,4-triazinone metamitron can be degraded to phenylglyoxalate by an Arthrobacter sp. (Figure 10.34a) (Engelhardt et al. 1982), whereas metribuzin undergoes mainly only transformation by reductive loss of the N-amino group and hydrolytic loss of the thiomethyl group (Figure 10.34b) (Lawrence et al. 1993).

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Nojiri H, J-W Nam, M Kosaka, K-I Morii, T Takemura, K Furihata, H Yamane, T Omori (1999) Diverse oxygenations catalyzed by carbazole 1,9a-dioxygenase from Pseudomonas sp. strain CA 10. J Bacteriol 181: 3105–3113. Nyns EJ, D Zach, EE Snell (1969) The bacterial oxidation of vitamin B6. VIII. Enzymatic breakdown of α-(N-acetylaminomethylene)succinic acid. J Biol Chem 244: 2601–2605. Ogawa J, S Shimizu (1995) Purification and characterization of dihydroorotase from Pseudomonas putida. Arch Microbiol 164: 353–357. Orpin CG, M Knight, WC Evans (1972) The bacterial oxidation of N-methylisonicotinate, a photolytic product of paraquat. Biochem J 127: 833–844. Padden AN, FA Rainey, DP Kelly, AP Wood (1997) Xanthobacter tagetidis sp. nov., an organism associated with Tagetes species and able to grow on substituted thiophenes. Int J Syst Bacteriol 47: 394–401. Parisot D, MC Malet-Martino, R Martino, P Crasnier (1991) 19F nuclear magnetic resonance analysis of 5-fluorouracil metabolism in four differently pigmented strains of Nectria haematococca. Appl Environ Microbiol 57: 3605–3612. Parshikov IA, JP Freeman, JO Lay, RD Beger, AJ Williams, JB Sutherland (2000) Microbiological transformation of enrofloxacin by the fungus Mucor ramannianus. Appl Environ Microbiol 66: 2664–2667. Peschke B, F Lingens (1991) Microbial metabolism of quinoline and related compounds XII. Isolation and characterization of the quinoline oxidoreductase from Rhodococcus spec. B1 compared with the quinoline oxidoreductase from Pseudomonas putida 86. Biol Chem Hoppe-Seyler 370: 1081–1088. Porter DJT (1994) Dehalogenating and NADPH-modifying activities of dihydropyrimidine dehydrogenase. J Biol Chem 269: 24177–24182. Rappert S, KC Botsch, S Nagorny, W Francke, R Müller (2006) Degradation of 2,3-diethyl-5-methylpyrazine by a newly discovered bacterium, Mycobacterium sp. Strain DM-11. Appl Environ Microbiol 72: 1437–1444. Resnick SM, DS Torok, DT Gibson (1993) Oxidation of carbazole to 3-hydroxycarbazole by naphthalene 1,2dioxygenase and biphenyl 2,3-dioxygenase. FEMS Microbiol Lett 113: 297–302. Rhee S-K, GM Lee, J-H Yoon, Y-H Park, H-S Bae, S-T Lee (1997) Anaerobic and aerobic degradation of pyridine by a newly isolated denitrifying bacterium. Appl Environ Microbiol 63: 2578–2585. Röger P, A Erben, F Lingens (1990) Microbial metabolism of quinoline and related compounds IV. Degradation of isoquinoline by Alcaligenes faecalis Pa and Pseudomonas diminuta 7. Biol Chem Hoppe-Seyler 370: 1183–1189. Röger P, G Bär, F Lingens (1995) Two novel metabolites in the degradation pathway of isoquinoline by Pseudomonas diminuta 7. FEMS Microbiol Lett 129: 281–286. Rosche B, B Shisuaka, S Getzner, F Lingens (1995) 2-oxo-1,2-dihydroquinoline 8-monooxygenase, a twocomponent enzyme system from Pseudomonas putida 86. J Biol Chem 270: 17836–17842. Sachelaru P, E Schiltz, GL Igloi, R Brandsch (2005) An α/β-fold C−C bond hydrolase is involved in a central step of nicotine metabolism by Arthrobacter nicotinovorans. J Bacteriol 187: 8516–8519. Sato S-I, J-W Nam, K Kasuga, H Nojiri, H Yamane, T Omori (1997b) Identification and characterization of genes encoding carbazole 1,9a-dioxygenase in Pseudomonas sp. strain CA10. J Bacteriol 179: 4850–4858. Sato, S-I, N Ouchiyama, T Kimura, H Nojiri, H Yamane, T Omori (1997a) Cloning of gees involved in carbazole degradation of Pseudomonas sp. strain CA10: nucleotide sequences of genes and characterization of meta cleavage enzymes and hydrolase. J Bacteriol 179: 4841–4849. Schach S, B Tshisuaka, S Fetzner, F Lingens (1995) Quinoline 2-oxidoreductase and 2-oxo-1,2-dihydroquinoline 5,6-dioxygenase from Comamonas testosteroni 63. The first two enzymes in quinoline and 3-methylquinoline degradation. Eur J Biochem 232: 536–544. Schenk S, A Hoelz, B Krauss, K Decker (1998) Gene structures and properties of enzymes of the plasmidencoded nicotine catabolism of Arthrobacter nicotinovorans. J Mol Biol 284: 1323–1339. Schiefer-Ullrich H, JR Andreesen (1985) Peptostreptoccus barnesae sp. nov, a Gram-positive, anaerobic, obligately purine utilizing coccus from chicken feces. Arch Microbiol 143: 26–31. Schieffer-Ullrich H, R Wagner, P Dürre, JR Andreesen (1984) Comparative studies on physiology and taxonomy of obligately purinolytic clostridia. Arch Microbiol 138: 345–353. Schmidt M, P Röger, F Lingens (1991) Microbial metabolism of quinoline and related compounds XI. Degradation of quinoline-4-carboxylic acid by Microbacterium sp. H2, Agrobacterium sp. 1B and Pimelobacter simplex 4B anf 5B. Biol Chem Hoppe-Seyler 370: 1015–1020. Schultz AC, P Nygaard, HH Saxild (2001) Functional analysis of 14 genes that constitute the purine catabolic pathway in Bacillus subtilis and evidence for a novel regulon controlled by the PucR transcription activator. J Bacteriol 183: 3293–3302.

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Watson GK, C Houghton, RB Cain (1974) Microbial metabolism of the pyridine ring. The hydroxylation of 4-hydroxypyridine to pyridine-3,4-diol (3,4-dihydroxypyridine) by 4-hydroxypyridine-3-hydroxylase. Biochem J 140: 265–276. Watson K, RB Cain (1975) Microbial metabolism of the pyridine ring. Metabolic pathways of pyridine biodegradation by soil bacteria. Biochem J 146: 157–172. Wetzstein H-G, M Stadler, H-V Tichy, A Dalhoff, W Karl (1999) Degradation of ciprofloxacin by basidiomycetes and identification of metabolites generated by the brown rot fungus Gloeophyllum striatum. Appl Environ Microbiol 65: 1556–1563. Wetzstein H-G, N Schmeer, W Karl (1997) Degradation of the fluoroquinolone enrofloxacin by the brown-rot fungus Gleophyllum striatum: identification of metabolites. Appl Environ Microbiol 63: 4272–4281. Wondrack LM, C-A Hsu, MT Abbott (1978) Thymine-7-hydroxylase and pyrimidine deoxyribonucleoside 2′-hydroxylase activities in Rhodotorula glutinis. J Biol Chem 253: 6511–6515. Woolfolk CA, BS Woolfolk, HR Whiteley (1970) 2-oxypurine dehydrogenase from Micrococcus aerogenes. I. Isolation, specificity, and some chemical and physical properties. J Biol Chem 245: 3167–3178. Wright KA, RB Cain (1972) Microbial metabolism of pyridinium compounds. Metabolism of 4-carboxy-1methylpyridinium chloride, a photolytic product of paraquat. Biochem J 128: 543–559. Xi H, BL Schneider, L Reitzer (2000) Purine catabolism in Escherichia coli and function of xanthine dehydrogenase in purine salvage J Bacteriol 182: 5332–5341. Xu G, TP West (1992) Reductive catabolism of pyrimidine bases by Pseudomonas steutzeri. J Gen Microbiol 138: 2459–2463. Xu G, TP West (1994) Characterization of dihydropyrimidase from Pseudomonas steutzeri. Arch Microbiol 161: 70–74.

PART 2 OXAARENES Naturally occurring oxaarenes based on polycyclic pyrans encompass a plethora of structures including the plant polyphenols such as anthocyanins and α-tocopherol (vitamin E). Halogenated dibenzo-p-dioxins and dibenzofurans are formed both as by-products during the manufacture of chlorophenols, and from the incineration of organic matter in the presence of inorganic halides.

AEROBIC CONDITIONS Monocyclic Oxaarenes The aerobic degradation of furan-2-carboxylate has already been summarized in Part 1 of this chapter and emphasis is placed here on anaerobic pathways. The degradation of furan-2-carboxaldehyde (furfural) has been examined in Desulfovibrio furfuralis and is initiated by dehydrogenation to furan-2-carboxylate followed by hydroxylation. This is hydrolyzed to 2-ketoglutarate and decarboxylated to succinic acid semialdehyde that is degraded to acetate (Figure 10.35a) (Folkerts et al. 1989). Furan-3-carboxylate is degraded by Paracoccus denitrificans strain MK33 under denitrifying conditions by the formation of the CoA ester and reduction at ∆4,5 before hydration (Figure 10.35b) (Koenig and Andreesen 1991). This strain is also able to degrade furan-3-carboxylate by an analogous pathway. Polycyclic Oxaarenes Monooxygenation Monooxygenation mediated by the cytochrome P450 system of Streptomyces has been observed in oxaarenes: a. 2,2-Dimethyl-6,7-dimethoxychromene was transformed to an epoxide, which then underwent further reactions (Figure 10.36) (Sariaslani et al. 1987). b. 7-Ethoxycoumarin was transformed initially to the 7-hydroxycompound, which was subsequently hydroxylated and partially O-methylated (Sariaslani et al. 1989).

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O

O

CO 2 H

CO 2 H

CO 2 H

CO 2 H

O

HO CO 2 H

CoSCoA

Acetate

O

O

O

CoSCoA

O

CO 2 H

O

O

e MeO

MeO

O

ne

ne

MeO

MeO

MeO

FIGURE 10.36 Transformation products from 2,2-dimethyl-7,8-dimethoxychromene.

MeO

MeO

O

MeO

O

OH

O

OH

CoSCoA

OH

OH

OH

+

e

MeO

MeO

MeO

MeO

HOCH 2

CO 2 H

HO 2 C

FIGURE 10.35 Anaerobic degradation of (a) furan-2-carboxylate and (b) furan-3-carboxylate.

(b)

(a)

CHO

O

OH

O

CoSCoA

OH

OH

CO 2 H

CO 2 H

HO 2 CCOCH 2 CH 2 CO 2 H

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555

Dioxygenation It is appropriate to note briefly the degradation of the analogous halogenated diphenyl ethers, which have been used as agrochemicals and flame-retardants. In Pseudomonas cepacia, the degradation of diphenyl ether was initiated by dioxygenation and produced pyr-2-one-6-carboxylate as the end product (Figure 10.37) (Pfeifer et al. 1993). A strain of Sphingomonas strain SS33 was able to degrade diphenyl ether and grow with 4-fluoro- and 4-chlorodiphenyl ether by dioxygenation to form 4-halogenated catechol and phenol (Schmidt et al. 1992). There are two variants for dioxygenation of polycyclic oxaarenes. Dioxygenation of the benzenoid ring occurs most commonly in transformations, rather than in degradations that proceed by angular dioxygenation at the ring junction in a range of polycyclic oxaarenes. For example, benzofuran is oxygenated by Pseudomonas putida strain UV4 to cis-(7S,6S)-benzofurandi-hydrodiol (the configuration at C7 is determined from the Sequence Rule that gives priority to 7S) (Boyd et al. 1992), dibenzofuran by Beijerinckia sp. strain B8/36 to the cis-2,3-dihydrodiol (Cerniglia et al. 1979), and dibenzo[1,4]dioxin by Pseudomonas sp. strain NCIB 9816 to the cis-1,2-dihydrodiol, which was dehydrogenated to 1,2dihydroxydibenzo[1,4]dioxin by naphthalene-grown cells (Klecka and Gibson 1979). The dioxygenases that carry out angular dioxygenation have been characterized in a number of degradations, which include not only dibenzofuran, dibenzo[1,4]dioxin, and xanthone, but also carbazole. All of them involve angular dioxygenation between the heteroatom and the benzene ring. The properties of the three-component dioxygenases have been summarized (Armengaud et al. 1998), and the three-component carbazole 1,9a dioxygenase from Pseudomonas resinovorans strain CA10 has been described (Nam et al. 2002). It has a relaxed specificity that includes dioxygenation not only of these heterocyclic compounds but also of the carbocyclic naphthalene, biphenyl, anthracene, and fluoranthene (Nojiri et al. 1999). The 3,4-dihydroxyxanthone dioxygenase (Chen and Tomasek 1991) that is used for the degradation of xanthone by a strain of Arthrobacter sp. (Figure 10.38) (Tomasek and Crawford 1986; Chen et al. 1986; Chen and Tomasek 1991) and the dioxygenase that carries out angular dioxygenation of dibenzofuran by a strain of Sphingomonas sp. have been characterized (Bünz and Cook 1993). A summary of the reactions involved in the biodegradation of dibenzo[1,4]dioxins, dibenzofurans, diphenyl ethers, and fluoren-9-one has been given (Wittich 1998). The degradation of 3-hydroxybenzofuran and 2-hydroxydibenzo[1,4]dioxin has been examined in Sphingomonas sp. strain RW1 and also involves angular dioxygenation. For both substrates, 1,2,4-trihydroxybenzene is produced from 3-hydroxybenzofuran by oxidative decarboxylation of 4-hydroxysalicylate (Armengaud et al. 1999). Xanthone

The degradation of xanthone by Arthrobacter sp. strain GFB100 closely resembled the pathways used for carbocyclic compounds. In a yeast-supplemented medium, initial dioxygenation to the 3,4dihydrodiol was followed by dehydrogenation to 3,4-dihydroxyxanthone and extradiol ring fission to a coumarin and subsequent production of 2,5-dihydroxybenzoate (Figure 10.38) (Chen and Tomasek 1991; Tomasek and Crawford 1986).

O

O OH

OH OH

CO 2 H

+ O

FIGURE 10.37

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O

OH

O

O

CO 2 H

Biodegradation of diphenyl ether.

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Environmental Degradation and Transformation of Organic Chemicals O

O

O

O

O

OH O

H

OH

OH H

OH

COCH 3 OH

O

OH

OH

CO 2 H O

O

O

O HO

CO 2 H

OH

FIGURE 10.38 Biodegradation of xanthone. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3J, pp. 1–80, Springer, Heidelberg, 1998. With permission.)

Dibenzofuran

The degradation of dibenzofuran by Pseudomonas sp. strain HH69 was initiated by dioxygenation with the formation of 2,2′,3-trihydroxybiphenyl that was further degraded to salicylate and several chroman-4-ones (Figure 10.39) (Fortnagel et al. 1990). The same strain (now designated as Sphingomonas sp.) is able to metabolize a range of substituted dibenzofurans (Harms et al. 1995). Other organisms that carry out the same initial reactions include Brevibacterium sp. strain DPO1361 (Strubel et al. 1991) and Sphingomonas sp. strain RW1 from which the 4,4a-dioxygenase has been isolated and characterized (Bünz and Cook 1993). Like the dioxygenases from carbocyclic compounds, the latter consists of four proteins, a reductase containing two isofunctional flavoproteins containing FAD that accepts electrons from NADH, a ferredoxin, and a terminal oxygenase containing a Rieske [2Fe–2S] center. It was suggested that the system involves a IIA dioxygenase. The dioxygenase genes in Terrabacter sp. strain YK3 are carried on a plasmid, and a phylogenetic analysis has compared the amino acid sequences of the gene products with those from a number of other dioxygenases (Iida et al. 2002). The key enzyme for the ring fission of 2,2′, 3-trihydroxybiphenyl has been biochemically and genetically analyzed in a strain of Sphingomonas sp. (Happe et al. 1993). Three extradiol dioxygenases are involved in the degradation of dibenzofuran by Terrabacter strain DPO360 (Brevibacter sp. strain DPO360)—two in ring fission of the intermediate 2,2′,3-trihydroxybiphenyl (BphC1 and BphC2) and a catechol 2,3-dioxygenase (Schmid et al. 1997). Dibenzo[1,4]dioxin

The analogous 4,4a-dioxygenation of dibenzo[1,4]dioxin by Sphingomonas sp. strain RW1 produced 2,2′,3-trihydroxydiphenyl ether that was degraded by an extradiol dioxygenase (Figure 10.40) (Happe et al. 1993) to catechol that underwent both intra- and extradiol fission (Wittich et al. 1992). A strain of Sphingomonas sp. grown with diphenyl ether was able to oxidize dibenzo[1,4]dioxin, and the versatility of members of the genus has been associated with the scattering of the genes for the component proteins of the dioxygenase system around the genome of Sphingomomas sp. strain RW1, which can degrade a number of dibenzo[1,4]dioxins (Armengaud et al. 1998).

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Heterocyclic Aromatic Compounds

557 OH

O

O OH OH

OH

OH

O CO 2 H OH

OH

O CO 2 H

O

CH 2−CO−CO2H

O

O

O

O

OH

CH 3

FIGURE 10.39 Pathways for degradation of dibenzofuran.

OH

O

O

O

HO O

O

H

OH

HO OH

O

OH

OH O

CO 2 H

OH

OH

FIGURE 10.40 Degradation of dibenzo[1,4]dioxin. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3J, pp. 1–80, Springer, Heidelberg, 1998. With permission.)

Plausible pathways for the degradation of 3-hydroxybenzofuran and 2-hydroxybenzo[1,4]dioxin have been proposed (Armengaud et al. 1999). Chlorinated Bibenzofuran and Dibenzo-p-Dioxin

Analogously, dioxygenation of these takes place at the angular positions, and in Sphingomonas sp. strain RW1 this is followed by ring fission that results in the formation of 4,5-dichlorosalicylate

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O

O

O

O

O

OH OH

OH

OH O OH CO 2 H

O

CO 2 H

OH

HO OH

OH

FIGURE 10.41 Degradation of 2,3-dichlorodibenzo[1,4]dioxin. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3J, pp. 1–80, Springer, Heidelberg, 1998. With permission.)

from 2,3-dichlorodibenzofuran and 6-chloro-2-methylchromenone from 2,8-dichlorodibenzofuran (Wilkes et al. 1996). Fission of the alternative ether groups in 2,3-dichlorodibenzo[1,4]dioxin produced 4,5-dichlorocatechol (Figure 10.41) (Habe et al. 2001; Hong et al. 2002), and analogously 3,4,5-trichlorocatechol from 1,2,3-trichlorodibenzo[1,4]dioxin, and 3,4,5,6-tetrachlorocatechol from 1,2,3,4,7,8-hexachlorodibenzo[1,4]dioxin by dioxygenation and fission of the ring with fewer chlorine substituents (Nam et al. 2006). Partial anaerobic dechlorination of chlorinated dibenzo[1,4]dioxins has been observed in sediment slurries. The 1,2,3,4-tetrachloro compound produced predominantly 1,3-dichlorodibenzo[1,4]dioxin (Beurskens et al. 1995; Ballerstedt et al. 1997), and for substrates with five to seven chlorine substituents, chlorine was removed from both the peri and the lateral positions (Barkovskii and Adriaens 1996).

Fungal Reactions The imperfect fungus Paecilomyces liliacinus is able to produce successively mono-, di-, and trihydroxylated metabolites from biphenyl, and carry out intradiol ring fission of rings carrying adjacent hydroxyl groups (Gesell et al. 2001). Analogously, mono- and dihydroxylated metabolites are produced from dibenzofuran followed by intradiol fission of the ring with adjacent hydroxyl groups (Gesell et al. 2004). Fungal peroxidases, which have already been noted in the degradation of 2,4dichlorophenol (Valli and Gold 1991) and 2,4-nitrotoluene (Valli et al. 1992a), are able to degrade 2,7-dichlorodibenzo[1,4]dioxin by a pathway completely different from those used by bacteria (Valli et al. 1992b) (Figure 10.42). Flavanoids, Isoflavanoids, and Related Compounds Many higher plants synthesize flavanes, flavanones, flavones, and isoflavones with a wide range of structural complexity. They make a significant contribution to the food intake of both herbivores and humans, and they have aroused particular interest on account of their degradation by mammals that are mediated by intestinal bacteria. Most of them exist naturally as glycosides and these are readily hydrolyzed to the aglycones.

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Heterocyclic Aromatic Compounds

559 OH

O−CH 3 OH

Cl O

Cl

O

O−CH 3

O−CH 3

Cl

O

Cl

O

O

OH OH

OH

O

CO 2 H CO 2 H

OH

O

FIGURE 10.42 Fungal degradation of 2,7-dichlorodibenzo[1,4]dioxin. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3J, pp. 1–80, Springer, Heidelberg, 1998. With permission.)

OH

OH OH HO

O

OH HO

O O CO 2 H

OH OH

FIGURE 10.43

O

+ CO

OH

Transformation of rutin by dioxygenation.

Reactions Carried Out by Fungi 1. The unusual dioxygenation of rutin by Aspergillus flavus to carbon monoxide and a depside of 3,4-dihydroxybenzoate and 2,4,6-trihydroxybenzoate (Figure 10.43) (Krishnamurty et al. 1970) is an unusual example of a dioxygenase synthesized by a eukaryotic microorganism (Krishnamurty and Simpson 1970). Experiments with 18O2 showed that both atoms were incorporated into the depside, whereas with H218O no incorporation occurred into either the depside or into carbon monoxide. The enzyme in Aspergillus japonicus that brings about fission of ring C with the formation of carbon monoxide is an unusual copper containing dioxygenase (Steiner et al. 2002). 2. The transformation of representatives of flavanones and isoflavanones has been examined in a number of fungi and in Streptomyces fulvissimus (Ibrahim and Abul-Hajj 1990a). The principal reactions involved dehydrogenation, aromatic ring hydroxylation, and reductive ring scission of flavanones (Ibrahim and Abul-Hajj 1990b), although the mechanisms of these transformations remain unresolved. These reactions are illustrated by an overview of the transformation of flavanone by Aspergillus niger NRRL 599 (Figure 10.44a) and of isoflavanone by A. niger X172 (Figure 10.44b).

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Environmental Degradation and Transformation of Organic Chemicals

O

O

OH

(a)

+ OH

HO

O

O

O

O

O

O

(b)

OH

+ O

O

O OH OH

FIGURE 10.44

Transformation of (a) flavanone and (b) isoflavanone.

Reactions Carried Out by Bacteria Aerobic Conditions

The key reaction under both aerobic and anaerobic conditions is hydrolysis of the 1,3-dicarbonyl intermediate involving the keto form of phloroglucinol or resorcinol. In species of Rhizobium and Brachyrhizobium, the aerobic degradation of flavanoids (naringenin, quercitin, and luteolin) is initiated by reductive fission of the C ring between C-2 and O followed by hydrolysis (Figure 10.45a) (Rao et al. 1991; Rao and Cooper 1994). In contrast, in the isoflavanoids (genistein and daidzein), reduction is accompanied by migration of the phenolic group from C-3 to C-2 (Figure 10.45b). An analogous pathway was followed in the degradation of quercitin by Pseudomonas putida strain PML2. It was initiated by reductive loss of hydroxyl groups at C-3 and C-3′ followed by fission of the C-ring to form phloroglucinol and 3,4-dihydroxycinnamic acid that was degraded to 3,4-dihydroxybenzoate before further degradation (Pillai and Swarup 2002). The dioxygenation of 7,4′-dihydroxyisoflavone by biphenyl-2,3-dioxygenase from Burkholderia sp. strain LB400 was unusual, since it is followed by elimination to the phenol 7,2′,4′-trihydroxyisoflavone rather than by dehydrogenation (Seeger et al. 2003). Anaerobic Conditions

The anaerobic degradation of flavanoids such as quercitin, naringin, and luteolin has been examined in a number of anaerobic bacteria including Butyrovibrio sp. (Cheng et al. 1971), Clostridium sp. (Winter et al. 1989), Eubacterium ramulus (Braune et al. 2001), and Clostridium orbiscindens (Schoefer et al. 2003). The initial reactions are enone reductions of the B ring followed by fission of the C2−O bond and hydrolysis (Figure 10.46a). For isoflavanoids such as daidzein two reactions after reduction have been found (i) fission of the C2−O bond in Clostridium sp. (Hur et al. 2002) (Figure 10.46b) or (ii) reduction of the ketone followed by dehydration to dehydroequol and a stereospecific reduction to S-equol (Figure 10.46c) (Wang et al. 2005).

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HO

OH

OH

O

O

O

O

OH

OH

OH HO

HO

OH

OH

O

OH

O

OH

OH

OH OH

OH

HO

HO

OH

OH

OH HO 2 C

CO 2 H

OH

+

HO 2 C

+

OH OH

OH

FIGURE 10.45 Aerobic transformation of (a) quercitin and (b) genistein. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3J, pp. 1–80, Springer, Heidelberg, 1998. With permission.)

(b)

(a)

HO

OH

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Environmental Degradation and Transformation of Organic Chemicals OH

OH

OH

OH HO

O

OH

(a)

HO

O

O

OH

O

OH OH HO

HO

OH

OH

HO

OH

+ OH

O

OH CO 2 H

HO

O

HO

O

O

O OH

HO

OH

(b)

(c)

CH 3 HO

O

O OH OH

HO

OH

O

OH

FIGURE 10.46 Anaerobic transformation of (a) quercitin and (b,c) daidzein. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3J, pp. 1–80, Springer, Heidelberg, 1998. With permission.)

REFERENCES Armengaud J, B Happe, KN Timmis (1998) Genetic analysis of dioxin dioxygenase of Sphingomonas sp. strain RW1: catabolic genes dispersed on the genome. J Bacteriol 180: 3954–3966. Armengaud J, KN Timmis, R-M Wittich (1999) A functional 4-hydroxysalicylate/hydroxyquinol degradative pathway gene cluster is linked to the initial dibenzo-p-dioxin pathway genes in Sphingomonas sp. strain RW1. J Bacteriol 181: 3452–3461. Ballerstedt H, A Kraus, U Lechner (1997) Reductive dechlorination of 1,2,3,4-tetrachlorodibenzo-p-dioxin and its products by anaerobic mixed cultures from Saale River sediment. Environ Sci Technol 31: 1749–1753.

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Barkovskii AI, P Adriaens (1996) Microbial dechlorination of historically present and freshly spiked chlorinated dioxins and diversity of dioxin-dechlorinating populations Appl Environ Microbiol 62: 4556–4562. Beurskens JEM, M Toussaint, J de Wolf, JMD van der Steen, PC Slot, LCM Commandeur, JR Parsons (1995) Dehalogenation of chlorinated dioxins by an anaerobic consortium from sediment. Environ Toxicol Chem 14: 939–943. Boyd DR, ND Sharma, R Boyle, RAS McMordie, J Chima, H Dalton (1992) A 1H NMR method for the determination of enantiomeric excess and absolute configuration of cis-dihydrodiol metabolites of polycyclic arenes and heteroarenes. Tetrahedron Lett 33: 1241–1244. Braune A, M Gütschow, W Engst, M Baut (2001) Degradation of quercitin and luteolin by Eubacterium ramulus. Appl Environ Microbiol 67: 558–5567. Bünz PV, AM Cook (1993) Dibenzofuran 4,4a-dioxygenase from Sphingomonas sp. strain RW1: angular dioxygenation by a three-component system. J Bacteriol 175: 6467–6475. Cerniglia CE, JC Morgan, DT Gibson (1979) Bacterial and fungal oxidation of dibenzofuran. Biochem J 180: 175–185. Chen C-M, PH Tomasek (1991) 3,4-Dihydroxyxanthone dioxygenase from Arthrobacter sp. strain GFB 100. Appl Environ Microbiol 57: 2217–2222. Chen C-M, PH Tomasek, RL Crawford (1986) Initial reactions of xanthone biodegradation by an Arthrobacter sp. J Bacteriol 167: 818–827. Cheng K-J, HG Krishnamurty, GA Jones, FJ Simpson (1971) Identification of products produced by the anaerobic degradation of naringin by Butyrivibrio sp. C3. Can J Microbiol 17: 129–131. Folkerts M, U Ney, H Kneifel, E Stackenbrandt, EG Witte, H Förstel, SM Schoberth, H Sahm (1989) Desulfovibrio furfuralis sp. nov., a furfural degrading strictly anaerobic bacterium. Syst Appl Microbiol 11: 161–169. Fortnagel P, H Harms, R-M Wittich, S Krohn, H Meter, V Sinnwell, H Wilkes, W Francke (1990) Metabolism of dibenzofuran by Pseudomonas sp. strain HH 69 and the mixed culture HH27. Appl Environ Microbiol 56: 1148–1156. Gesell M, E Hammer, A Mikolasch (2004) Oxidation and ring cleavage of dibenzofuran by the filamentous fungus Paecilomyces lilacinus. Arch Microbiol 182: 51–59. Gesell M, E Hammer, M Specht, W Francke, F Schauer (2001) Biotransformation of biphenyl by Paecilomyces lilacinus and characterization of ring cleavage products. Appl Environ Microbiol 67: 1551–1557. Habe H, J-S Chung, J-H Lee, K Kasuga, T Yoshida, H Nojiri, T Omori (2001) Degradation of chlorinated dibenzofurans and dibenzo-p-dioxins by two types of bacteria having angular dioxygenases with different features. Appl Environ Microbiol 67: 3610–3617. Happe B, LD Eltis, H Poth, R Hedderich, KN Timmis (1993) Characterization of 2,2′,3-trihydroxybiphenyl dioxygenase, an extradiol dioxygenase from the dibenzofuran- and dibenzo-p-dioxin-degrading bacterium Sphingomonas sp. strain RW1. J Bacteriol 175: 7313–7320. Harms H, H Wilkes, R-M Wittich, P Fortnagel (1995) Metabolism of hydroxydibenzofurans, methoxydibenzofurans, acetoxydibenzofurans, and nitrodibenzofurans by Sphingomonas sp. strain HH69. Appl Environ Microbiol 61: 2499–2505. Hong H-B, Y-S Chang, I-H Nam, P Fortnagel, S Schmidt (2002) Biotransformation of 2,7-dichloro- and 1,2,3, 4-tetrachlorodibenzo-p-dioxin by Sphingomonas wittichii RW1. Appl Environ Microbiol 68: 2584–2588. Hur H-G, JO Lay, RD Beger, JP Freeman, F Rafii (2002) isolation of human intestinal bacteria metabolizing the natural isoflavone glycosides daidzin and genistin. Arch Microbiol 174: 422–428. Ibrahim A-R S, YJ Abul-Hajj (1990a) Microbiological transformation of flavone and isoflavone. Xenobiotica 20: 363–373. Ibrahim A-R S, YJ Abul-Hajj (1990b) Microbiological transformation of (+/−)-flavanone and (+/−)-isoflavanone. J Nat Prod 53: 644–656. Iida T, Y Mukouzaka, K Nakamura, T Kudo (2002) Plasmid-borne genes code for an angular dioxygenase involved in dibenzofuran degradation by Terrabacter sp. strain YK3. Appl Environ Microbiol 68: 3716–3723. Klecka GM, DT Gibson (1979) Metabolism of dibenzo[1,4]dioxan by a Pseudomonas species. Biochem J 180: 639–645. Koenig K, JR Andressen (1991) Aerobic and anaerobic degradation of furan-3-carboxylate by Paracoccus denitrificans strain MK 33. Arch Microbiol 157: 70–75. Krishnamurty HG, FJ Simpson (1970) Degradation of rutin by Aspergillus flavus. Studies with oxygen 18 on the action of a dioxygenase on quercitin. J Biol Chem 245: 1467–1471.

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Krishnamurty HG, K-J Cheng, GA Jones, FJ Simpson, JE Watkin (1970) Identification of products produced by the anaerobic degradation of rutin and related flavonoids by Butyrivibrio sp. C3. Can J Microbiol 16: 759–767. Nam I-H, Y-M Kim, S Schmidt, Y-S Chang (2006) Biotransformation of 1,2,3-tri- and 1,2,3,4,7,8-hexachlorodibenzo-p-dioxin by Sphingomonas wittichii strain RW1. Appl Environ Microbiol 72: 112–116. Nam J-W, H Nojiri, H Noguchi, H Uchimura, T Yoshida, H Habe, H Yamane, T Omori (2002) Purification and characterization of carbazole 1,9a-dioxygenase, a three-component dioxygenase system of Pseudomonas resinovorans strain CA10. Appl Environ Microbiol 68: 5882–5890. Nojiri H, J-W Nam, M Kosaka, K-I Morii, T Takemura, K Furihata, H Yamane, T Omori (1999) Diverse oxygenations catalyzed by carbazole 1,9a-dioxygenase from Pseudomonas sp. strain CA 10. J Bacteriol 181: 3105–3113. Pfeifer F, HG Trüper, J Klein, S Schacht (1993) Degradation of diphenylether by Pseudomonas cepacia Et4: enzymatic release of phenol from 2,3-dihydroxydiphenylether. Arch Microbiol 159: 323–329. Pillai BVS, Swarup (2002) Elucidation of the flavonoid catabolism pathway in Pseudomonas putida PML2 by comparative metabolic profiling. Appl Environ Microbiol 68: 143–151. Rao JR, E Cooper (1994) Rhizobia catabolize nod gene-inducing flavonoids via C-ring fission mechanisms. J Bacteriol 176: 5409–5413. Rao JR, ND Sharma, JTG Hamilton, DR Boyd, JE Cooper (1991) Biotransformation of the pentahydroxy flavone quercitin by Rhizobium loti and Bradyrhizobium strains (Lotus). Appl Environ Microbiol 57: 1563–1565. Sariaslani FS, LR McGee, DW Ovenall (1987) Microbial transformation of precocene II: oxidative reactions by Streptomyces griseus. Appl Environ Microbiol 53: 1780–1784. Sariaslani FS, MK Trower, SE Buchloz (1989) Xenobiotic transformations by Streptomyces griseus. Dev Ind Microbiol 30: 161–171. Schmid AS, B Rothe, J Altenbuchnerm W Ludwig, K-H Engesser (1997) Characterization of three distinct extradiol dioxygenases involved in mineralization of dibenzofuran by Terrabacter sp. strain DPO 360. J Bacteriol 179: 53–62. Schmidt S, R-M Wittich, D Erdmann, H Wilkes, W Francke, P Fortnagel (1992) Biodegradation of diphenyl ether and its monohalogenated derivatives by Sphingomonas sp. strain SS3. Appl Environ Microbiol 58: 2744–2750. Schoefer L, R Mohan, A Schwiertz, A Braune, M Blaut (2003) Anaerobic degradation of flavonoids by Clostridium orbiscindens. Appl Environ Microbiol 69: 5849–5854. Seeger M, M González, B Cámara, L Mu∼noz, E Ponce, L Mejías, C Mascayano, Y Vásquez, S SepúlvedaBoza (2003) Biotransformation of natural and synthetic isoflavanoids by two recombinant microbial enzymes. Appl Environ Microbiol 69: 5045–5050. Steiner RA, KH Kalk, BW Dijkstra (2002) Anaerobic enzyme.substrate structures provide insight into the reaction mechanism of the copper-dependent quercitin 2,3-dioxygenase. Proc Natl Acad USA 99: 16625–16630. Strubel V, K-H Engesser, P Fischer, H-J Knackmuss (1991) 3-(2-hydroxyphenyl)catechol as substrate for proximal meta ring cleavage in dibenzofuran degradation by Brevibacterium sp. strain DPO 1361. J Bacteriol 173: 1932–1937. Tomasek PH, RL Crawford (1986) Initial reactions of xanthone biodegradation by an Arthrobacter sp. J Bacteriol 167: 818–827. Valli K, BJ Brock, DK Joshi, MH Gold (1992a) Degradation of 2,4-dinitrotoluene by the lignin-degrading fungus Phanerochaete chrysosporium. Appl Environ Microbiol 58: 221–228. Valli K, H Warishi, MH Gold (1992b) Degradation of 2,7-dichlorodibenzo-p-dioxin by the lignin-degrading basidiomycete Phanerochaete chrysosporium. J Bacteriol 174: 2131–2137. Valli K, MH Gold (1991) Degradation of 2,4-dichlorophenol by the lignin-degrading fungus Phanerochaete chrysosporium. J Bacteriol 173: 345–352. Wang X-Y, H-G Hur, JH Lee, KT Kim, S-I Kim (2005) Enantioselective synthesis of S-equol from dihydrodaidzein by a newly isolated anaerobic human intestinal bacterium. Appl Environ Microbiol 71: 214–219. Wilkes H, RM Wittich, KN Timmis, P Fortnagel, W Francke (1996) Degradation of chlorinated dibenzofurans and dibenzo-p-dioxins by Sphingomonas sp. strain RW1. Appl Environ Microbiol 62: 367–371. Winter J, LH Moore, VR Dowell, VD Bokkenheuser (1989) C-ring cleavage of flavonoids by human intestinal bacteria. Appl Environ Microbiol 55: 1203–1208.

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Wittich R-M (1998) Degradation of dioxin-like compounds by microorganisms. Appl Microbiol Biotechnol 49: 489–499. Wittich, R-M, H Wilkes, V Sinnwell, W Francke, P Fortnagel (1992) Metabolism of dibenzo-p-dioxin by Sphingomonas sp. strain RW1. Appl Environ Microbiol 58: 1005–1010.

PART 3 THIAARENES: BENZOTHIOPHENES, DIBENZOTHIOPHENES, AND BENZOTHIAZOLE Thiaarenes are major components of crude oil and extensive effort has been devoted to microbial processes for their removal since their presence generates undesirable SOx during incineration of fossil fuels. Attention has been directed to a number of organisms including species of Rhodococcus, Corynebacterium, and Gordonia. Benzothiophene is isoelectronic with naphthalene, dibenzothiophene with anthracene, and benzothiazole with quinoline, and this is reflected in their aerobic degradation that is initiated by dioxygenation. The diversity of pathways for the degradation of dibenzothiophene is illustrated by the following examples: 1. By analogy with anthracene, dioxygenation of dibenzothiophene produces 2-hydroxybenzothiophene-3-carboxaldehyde, and this may be transformed into a number of products including benzothiophene-2,3-dione and further into disulfides and thioindigo (Bressler and Fedorak 2001a,b). 2. Dioxygenation of one of the rings occurs, and after dehydrogenation to a catechol, ring fission takes place (Figure 10.47) (Kodama et al. 1973). This pathway is analogous to that used for the degradation of naphthalene and the isoelectronic anthracene. 3. Considerable attention has been directed to reactions in which the sulfur is removed from dibenzothiophene: a. Successive oxidation at the sulfur atom may take place with the formation of the sulfoxide, followed by elimination of sulfite to yield either 2-hydroxybiphenyl (Omori et al. 1992; Rhee et al. 1998) or benzoate (Van Afferden et al. 1990) (Figure 10.48). b. Alternatively, when they serve as sources of sulfur, transformation with elimination of sulfite can be initiated in Rhodococcus sp. strain WU-K2R by oxidation at the sulfur atom to sulfones, followed by a complex series of reactions to produce 2-hydroxystyrene and benzofuran (Kirimura et al. 2002). c. In Rhodococcus erythropolis strain D-1, four flavin reductases designated DszC, DszB, and DszA bring about analogous oxygenations of dibenzothiophene to form 2′-hydroxybiphenyl sulfinate that loses sulfite to produce 2-hydroxybiphenyl (Figure 10.49) (Oldfield et al. 1997; Matsubara et al. 2001).

OH

H

OH

H

OH OH S

S OH

S OH

CO−CO 2H S

FIGURE 10.47

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CHO S

Alternative pathways for the biodegradation of dibenzothiophene.

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S

S

S O

O

O

CO 2 H + SO 3

2−

+ SO 3 2− OH

FIGURE 10.48

Alternative pathways for the biodegradation of dibenzothiophene.

S

S

S O

O

O

CO 2 H + SO 3

2−

+ SO 3

2−

OH

FIGURE 10.49

Degradation of benzothiophene to 2-hydroxybiphenyl. SO 2 H

S

S O

O

OH

CHO

OH

FIGURE 10.50 Degradation of benzothiophene.

d. In Rhodococus sp. strain ECRD-1, degradation involves successive oxidation to the sulfoxide that is further degraded primarily via 2′hydroxybiphenyl-2-sulfonate to 2-hydroxybiphenyl and sulfate (Macpherson et al. 1998): the cyclized sultones and sultines were also isolated. e. In Gordonia sp. strain 213E, successive oxidation of benzothiophene to the sulfoxide is followed by fission of the thiophene ring to 2-(2′hydroxyphenyl)ethen-1sulfinate that is converted into 2-(2′hydroxyphenyl)ethan-1-al (Gilbert et al. 1998) (Figure 10.50). f. Benzothiophene is oxidized by strains of Pseudomonas sp. to both sulfoxide and sulfone (Kropp et al. 1994a), and the sulfoxide undergoes an abiotic reaction with the formation of benzo[b]naphtho[1,2-d]thiophene (Figure 10.51) (Kropp et al. 1994b). 4. The Corynebacterium sp. that utilizes dibenzothiophene as a sulfur source produced 2-hydroxybiphenyl, and subsequently, nitrated this using the nitrate in the growth

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567

SO S

S

S

S

O

O

FIGURE 10.51 Transformation of benzothiophene involving microbial oxidation and chemical reaction of the sulfoxide. (From Neilson, A.H. and Allard, A.-S., The Handbook of Environmental Chemistry, Vol. 3J, pp. 1–80, Springer, Heidelberg, 1998. With permission.)

O2 N (a)

+ OH

OH

OH NO 2 HO

(b) Cl

OH + Cl

Cl NO 2

NO 2

FIGURE 10.52 Formation of nitrohydroxybiphenyls during metabolism of (a) dibenzothiophene and (b) 4-chlorobiphenyl.

medium to form two hydroxynitrobiphenyls (Omori et al. 1992) (Figure 10.52a). This reaction is reminiscent of a similar one that takes place during the metabolism of 4-chlorobiphenyl (Figure 10.52b) (Sylvestre et al. 1982). These products are plausibly formed from arene oxide intermediate produced by monooxygenase systems and are discussed in Chapter 2. Under anaerobic conditions dibenzothiophene is metabolized to a carboxybenzothiophene whose structure has not been established (Annweiler et al. 2001).

BENZOTHIAZOLE 2-Marcaptobenzothiazole is widely used as a rubber vulcanizer and as a corrosion inhibitor. The degradation of benzothiazole by two strains of Rhodococcus proceeds by formation of 2,6-dihydroxybenzothiazole (Besse et al. 2001). In contrast, degradation by a pyridine-degrading strain of Rhodococcus pyridinivorans proceeded by hydroxylation at C2, followed by the formation of the 6,7-diol and fission by both intradiol and extradiol catechol dioxygenases to a thiazolone dicarboxylate (Figure 10.53) (Haroune et al. 2002). Although biotransformation of 2-marcaptobenzothiazole by Rhodococcus rhodochrous may produce the S-methyl compound as a terminal metabolite, the aerobic biodegradation may also proceed by dioxygenation at the 6,7-positions followed by intradiol fission of the catechol to a mercaptothiazole dicarboxylic acid (Haroune et al. 2004).

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N

S

S

HO 2 C

N OH

N O

OH S

HO

HO 2 C

S

OH

FIGURE 10.53 Aerobic degradation of benzothiazole.

REFERENCES Annweiler E, W Michaelis, RU Meckenstock (2001) Anaerobic cometabolic conversion of benzothiophene by a sulfate-reducing enrichment culture and in a tar-oil-contaminated aquifer. Appl Environ Microbiol 67: 5077–5083. Besse P, B Combourieu, G Boyse, M Sancelme, H de Wever, A-M Delort (2001) Long-range 1H-15N heteronuclear shift correlation at natural abundance: a tool to study benzothiazole biodegradation by two Rhodococcus strains. Appl Environ Microbiol 67: 1412–1417. Bressler DC, PM Fedorak (2001a) Purification, stability, and mineralization of 3-hydroxy-2-formylbenzothiophene, a metabolite of dibenzothiophene. Appl Environ Microbiol 67: 821–826. Bressler DC, PM Fedorak (2001b) Identification of disulfides from the biodegradation of dibenzothiophene. Appl Environ Microbiol 67: 5084–5093. Gilbert SC, J Morton, S Buchanan, C Oldfield, A McRoberts (1998) Isolation of a unique benzothiophenedesulphurizing bacterium, Gordona sp. strain 213E (NCIMB 40816), and characterization of the desulphurization pathway. Microbiology (UK) 144: 2545–2553. Haroune N, B Combourieu, P Besse, M Sancelme, A Kloepfer, T Reemtsma, H De Wever, A-M Delort (2004) Metabolism of 2-mercaptobenzothiazole by Rhodococcus rhodochrous. Appl Environ Microbiol 70: 6315–6319. Haroune N, B Combourieu, P Besse, M Sancelme, T Reemtsma, A Kloepfer, A Diab, JS Knapp, S Baumberg, A-M Delort (2002) Benzothiazole degradation by Rhodococcus pyridinovorans strain PA: evidence of a catechol 1,2-dioxygenase activity. Appl Environ Microbiol 68: 6114–6120. Kirimura K, T Furuya, R Sato, Y Ishii, K Kino, S Usami (2002) Biodesulfurizarion of naphthothiophene and benzothiophene through selective cleavage of carbon–sulfur bonds by Rhodococcus sp. strain WUK2R. Appl Environ Microbiol 68: 3867–3872. Kodama K, K Umehara, K Shikmizu, S Nakatani, Y Minoda, K Yamada (1973) Identification of microbial products from dibenzothiophene and its proposed oxidation pathway. Agric Biol Chem 37: 45–50. Kropp KG, JA Goncalves, JT Anderson, PM Fedorak PM (1994a) Bacterial transformations of benzothiophene and methylbenzothiophenes. Environ Sci Technol 28: 1348–1356. Kropp KG, JA Goncalves, JT Anderson, PM Fedorak (1994b) Microbially mediated formation of benzonaphthothiophenes from benzo[b]thiophenes. Appl Environ Microbiol 60: 3624–3631. Macpherson T, CW Greer, E Zhou, AM Jones, G Wisse, PCK Lau, B Sankey, MJ Grossman, J Hawari (1998) Application of SPME/GC-MS to characterize metabolitres in the biodesulfurization of organosulur model compounds in bitumen. Environ Sci Technol 32: 421–426. Matsubara T, T Ohshiro, Y Nishina, Y Izumi (2001) Purification, characterization, and overexpression of flavin reductase involved in dibenzothiophene desulfurization by Rhodococcus erythropolis D-1. Appl Environ Microbiol 67: 1179–1184. Oldfield C, O Pogrebinsky, J Simmonds, ES Olson, CF Kulpa (1997) Elucidation of the metabolic pathway for dibenzothiophene desulphurization by Rhodococcus sp. strain IGTS8 (ATCC 53968). Microbiology (UK) 143: 2961–2973. Omori T, L Monna, Y Saiki, T Kodama (1992) Desulfurization of dibenzothiophene by Corynebacterium sp. strain SY1. Appl Environ Microbiol 58: 911–915. Rhee S-K, JH Chang, YK Chang, HN Chang (1998) Desulfurization of dibenzothiophene and diesel oils by a newly isolated Gordona strain CYKS1. Appl Environ Microbiol 64: 2327–2331. Sylvestre M, R Massé, F Messier, J Fauteux, J-G Bisaillon, R Beaudet (1982) Bacterial nitration of 4-chlorobiphenyl. Appl Environ Microbiol 44: 871–877. Van Afferden M, S Schacht, J Klein, HG Trüper (1990) Degradation of dibenzothiophene by Brevibacterium sp. DO. Arch Microbiol 153: 324–328.

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11

Miscellaneous Compounds

Parts 1 through 5 of this chapter provide an outline of the reactions involved in the biodegradation of aliphatic esters, ethers, nitramines, phosphonates and sulfonates, and organic compounds of metals and metalloids.

PART 1 CARBOXYLATE, SULFATE, PHOSPHATE, AND NITRATE ESTERS Hydrolysis is the first step in the degradation of a number of important contaminants. These include o-phthalate esters used as plasticizers; organophosphate, organothiophosphate, and organodithiophosphate insecticides; phosphorofluoridates that have been considered as chemical warfare agents; aliphatic sulfates in surfactants; and polyester polyurethane (Akutsu et al. 1998). In addition, there are important naturally occurring phosphonates, phosphates, and pyrophosphates; polysaccharide sulfates; linear alkyl (C16 –C30) sulfates; and sulfate conjugate esters of phenolic compounds such as tyrosine sulfate. The hydrolysis of a range of naturally occurring sulfate esters may make an important contribution to the sulfate present in aerobic soils (Fitzgerald 1976), quite apart from the anthropogenic contribution of SOx.

CARBOXYLATES Phthalate esters are widespread contaminants and the dialkyl phthalates are hydrolyzed before degradation of the resulting phthalate (Eaton and Ribbons 1982), which is discussed in Chapter 8, Part 3. For dimethyl phthalate, dimethyl terephthalate, and dimethyl isophthalate only partial hydrolysis may take place (Li et al. 2005). Cocaine is hydrolyzed to benzoate and ecgonine methyl ester by a strain of Pseudomonas maltophilia (Britt et al. 1992).

SULFATES Alkyl sulfates and alkylethoxy sulfates have been extensively used as detergents so that concern has been expressed over their biodegradability. A review (Cain 1981) covers the degradation of a wide range of surfactants including both of these groups, and one by White and Russell (1994) discusses in detail the biodegradation of alkyl sulfates including the enzymology and regulation. Although sulfate is formed by hydrolysis of both alkyl and aryl sulfates, the pathway of degradation for aryl sulfates is controlled by the source of sulfur (Cook et al. 1999). The complex issues surrounding the hydrolysis of sulfate esters have been discussed (Kertesz 1999), and are illustrated by the number and substrate specificity of alkyl sulfatases for the surfactant 2-butyloctyl sulfate in Pseudomonas sp. strain AE-A (Ellis et al. 2002). Long-chain unbranched aliphatic sulfate esters are generally degraded by initial hydrolysis to sulfate and the alkanol, which is then degraded by conventional pathways. The alkylsulfatases show diverse specificity (Dodson and White 1983)—generally for sulfates with chain lengths >5— although organisms have been isolated that degrade short-chain (C1–C4) primary alkyl sulfates (White et al. 1987). Pseudomonas sp. strain C12B produces two primary (P1 and P2) and three secondary (S1, S2, and S3) alkylsulfatases, and hydrolysis by the P1 enzyme takes place by O–S bond fission. The S1 and S2 enzymes are constitutive, and for chiral compounds such as octan-2-yl 569

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(a)

CH3

CH−O−SO3H2

CH3

(b)

FIGURE 11.1

CH3−O−SO3H2

CH3−CH−CO2H

CH3−CH(OH)−CO2H

O−SO3H2

HO−CH2−O−SO3H2

CH2O

Degradation of (a) propan-2-yl sulfate and (b) methyl sulfate.

sulfate, hydrolysis proceeds with inversion of configuration by cleavage of the alkyl–oxygen bond (Bartholomew et al. 1977). For example, hydrolysis of (R)-2-octyl sulfate by Rhodococcus ruber DSM 44541 proceeded with inversion to (S)-2-octanol (Pogorevc and Faber 2003). As an alternative pathway for degradation, oxidation may precede elimination of sulfate. Examples include the degradation of propan-2-yl sulfate (Crescenzi et al. 1985) and of monomethyl sulfate (Davies et al. 1990; Higgins et al. 1993) (Figure 11.1). Under sulfate limitation, Pseudomonas putida strain S-313 degraded alkyl sulfates (C4 –C12) to sulfate and the corresponding aldehyde by a 2-ketoaciddependent reaction in the presence of Fe2+. Exceptionally, other 2-ketoacids could be used including 2-ketoglutarate, 2-ketovalerate, and 2-ketoadipate (Kahnert and Kertesz 2000). For alkylethoxy sulfates, a greater range of possibilities exist including ether-cleavage reactions, while direct removal of sulfate may be of lesser significance (Hales et al. 1986). It should be noted, however, that an unusual reaction may occur simultaneously: chain elongation of the carboxylic acid. For example, during degradation of dodecyl sulfate, lipids containing 14, 16, and 18 carbon atoms were synthesized (Thomas and White 1989). Transformations by monooxygenation rather than hydrolysis have been described for the organochlorine insecticides endosulfan and endosulfate in organisms that use them as sources of sulfur in the absence of inorganic sulfur. Hydroxylation at the position adjacent to the oxygen is mediated by a monooxygenase and a reduced flavin, and results in the production of sulfite from β-endosulfan by Mycobacterium sp. strain ESD (Sutherland et al. 2002). Analogous monooxygenation of α- and β-endosulfan, and endosulfate can be carried out by a species of Arthrobacter (Weir et al. 2006). Aryl sulfates are widely synthesized from phenolic substrates and serve as a detoxification mechanism both for microorganisms and for fish. Hydrolysis of aryl esters takes place with fission of the O–S bond (Recksiek et al. 1998), whereas for aliphatic sulfates, fission of the C–O bond takes place. In addition, utilization of aryl sulfate may be mediated by an arylsulfotransferase (Kahnert et al. 2000). As for alkyl sulfates, sulfate is formed from aryl sulfates and, for aryl sulfates the pathway of degradation is controlled by the source of sulfur (Cook et al. 1999). The hydrolysis of aryl sulfates has traditionally been a useful taxonomic character in the genus Mycobacterium (Wayne et al. 1991). A positive result in the 3-d aryl sulfatase test based on the hydrolysis of phenolphthalein sulfate has been particularly valuable for distinguishing members of the rapidly growing M. fortuitum group, which are potentially pathogenic to man, whereas the slow-growing M. tuberculosis and M. bovis are negative even after 10 d. Regulation of the synthesis of tyrosine sulfate sulfohydrolase has been examined in a strain of Comamonas terrigena (Fitzgerald et al. 1979), and both inducible and constitutive forms of the enzyme exist. These are, however, apparently distinct from the aryl sulfate sulfohydrolase, which has found application in taxonomic classification.

PHOSPHATES Aryl phosphates and thiophosphates, and alkyl dithiophosphates are important agrochemicals, while phosphorofluoridates have been prepared as chemical warfare agents. Concern over the persistence and the biodegradability of organophosphate and organophosphorothioates, which are used as agrochemicals, has stimulated studies into their degradation. Considerable attention has been

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directed to biodegradation of all of them, and references may be found in Munnecke et al. (1982), DeFrank and White (2002), and Singh et al. (2004). The hydrolytic enzyme(s)—organophosphorus acid anhydrase (OPA)—responsible for defluorination of phosphofluoridates are noted in Chapter 14, Part 6. These studies have revealed the widespread distribution of bacterial triesterases, whereas diesterases are less common, although an enzyme from Delftia acidovorans has been described (Tehara and Keasling 2003). The first step in the degradation of phosphate and phosphorothioate esters is hydrolysis, and substantial effort has been directed to all groups. Investigations have also been directed to the use of their degradation products as a source of phosphate for the growth of bacteria, and a wide range of phosphates, dialkylphosphates, and phosphorothioates has therefore been examined as sources of phosphorus (Cook et al. 1978). It is important to emphasize that the initial metabolites after hydrolysis may be both toxic and sometimes resistant to further degradation. Examples include nitrophenols, whose degradation is discussed in Chapter 9, Part 5; and 3,5,6-trichloropyridin-2-ol (Feng et al. 1997), which is produced by the hydrolysis of chlorpyrifos (O,O-diethyl-O-[3,5,6-trichlo-2-pyridyl]phosphorothioate).

NITRATES Compared with the fairly numerous investigations on the microbial degradation of carboxylic acid, sulfate, and phosphate esters, data on the degradation of nitrate esters are more fragmentary. This has been clearly revealed in a review (White and Snape 1993), which summarized existing knowledge on the microbial degradation of nitrate esters including glycerol trinitrate and its close relatives, and the pharmaceutical products pentaerythritol tetranitrate and isosorbide 2,5-dinitrate. Examples include glycerol trinitrate (Marshall and White 2001) and pentaerythritol tetranitrate (French et al. 1996), which are important explosives. Reactions involving glutathione transferases are important in eukaryotic microorganisms (White et al. 1996). In bacteria, however, the degradation of nitrate esters takes place by reduction with loss of nitrite (Figure 11.2), and pentaerythritol tetranitrate reductase, which is related to “old yellow enzyme” (French et al. 1996), and glycerol trinitrate reductase (Snape et al. 1997) have been purified. The reductase from E. cloacae is strongly inhibited by steroids and is capable of the reduction of cyclohex-2-ene-1-one (French et al. 1996). The biotransformation of gylcerol trinitrate by strains of Bacillus thuringiensis/cereus or Enterobacter agglomerans (Meng et al. 1995), by strains of Pseudomonas sp., and some Enterobacteriaceae (Blehert et al. 1997) involves the expected successive loss of nitrite with the formation of glycerol. The biotransformation of pentaerythritol tetranitrate by Enterobacter cloacae proceeds comparably with metabolism of two hydroxymethyl groups produced by loss of nitrite to the aldehyde (Binks et al. 1996). In a medium containing glucose and ammonium nitrate, glyceryl trinitrate is degraded by Penicillium corylophilum to the di- and mononitrate before complete degradation. In contrast, the metabolism of glyceryl trinitrate by Phanerochaete chrysosporium involves the production of nitric oxide (Servent et al. 1991). Nitric oxide is produced during conversion of l-arginine into l-citrulline by a strain of Nocardia sp. (Chen and Rosazza 1995), and a summary of the mechanism, which includes both prokaryotic and mammalian systems, has been given (Stuehr et al. 2004).

O2NOCH2 O2NOCH2

FIGURE 11.2

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C

CH2O−NO2 CH2−O−NO2

O2NOCH2 HOCH2

C

CH2OH CH2−O−NO2

Reductive degradation of pentaerythritol tetranitrate.

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REFERENCES Akutsu Y, T Nakajima-Kambe, N Nomura, T Nakahara (1998) Purification and properties of a polyester polyurethane-degrading enzyme from Comamonas acidovorans TB-35. Appl Environ Microbiol 64: 62–67. Bartholomew B, KS Dodgson, GWJ Matcham, DJ Shaw, GF White (1977) A novel mechanism of enzymatic hydrolysis. Inversion of configuration and carbon–oxygen bond cleavage by secondary alkylsulphohydrolases from detergent-degrading micro-organisms. Biochem J 165: 575–580. Binks PR, CE French, S Nicklin, NC Bruce (1996) Degradation of pentaerythritol tetranitrate by Enterobacter cloacae PB2. Appl Environ Microbiol 62: 1214–1219. Blehert DS, KL Knoke, BG Fox, GH Cambliss (1997) Regioselectivity of nitroglycerine denitration by flavoprotein nitroester reductases purified from two Pseudomonas species. J Bacteriol 179: 6912–6920. Britt AJ, NC Bruce, CR Lowe (1992) Identification of a cocaine esterase in a strain of Pseudomonas maltophilia. J Bacteriol 174: 2087–2094. Cain RB (1981) Microbial degradation of surfactants and “builder” components. In Microbial Degradation of Xenobiotics and Recalcitrant Compounds (Eds T Leisinger, AM Cook, R Hütter, and J Nüesch), pp. 325–370. Academic Press, London. Chen Y, JPN Rosazza (1995) Purification and characterization of nitric oxide synthase NOSNoc from a Nocardia sp. J Bacteriol 177: 5122–5128. Cook AM, CG Daughton, M Alexander (1978) Phosphorus-containing pesticide breakdown products: quantitative utilization as phosphorus sources by bacteria. Appl Environ Microbiol 36: 668–672. Cook AM, H Laue, F Junker (1999) Microbial desulfonation FEMS Microbiol Rev 22: 399–419. Crescenzi AMV, KS Dodgson, GF White, WJ Payne (1985) Initial oxidation and subsequent desulphation of propan-2-yl sulphate by Pseudomonas syringae strain GG. J Gen Microbiol 131: 469–477. Davies I, GF White, WJ Payne (1990) Oxygen-dependent desulphation of monomethyl sulphate by Agrobacterium sp. M3C. Biodegradation 1: 229–241. DeFrank JJ, WE White (2002) Phosphorofluoridates: biological activity and biodegradation. Handbook Environ Chem 3N: 295–343. Dodson KS, GF White (1983) Some microbial enzymes involved in the biodegradation of sulfated surfactants. In Topics in Enzyme and Fermentation Technology (Ed A Wiseman), Vol 7, pp. 90–155. Ellis-Horwood, Chichester. Eaton RW, DW Ribbons (1982) Metabolism of dibutylphthalate and phthalate by Micrococcus sp. strain 12B. J Bacteriol 151: 48–57. Ellis AW, SG Hales, NGA Ur-Rehman, GF White (2002) Novel alkylsulfatases required for biodegradation of the branched primary alkyl sulfate surfactant 2-butyloctyl sulfate. Appl Environ Microbiol 68: 31–36. Feng Y, KD Racke, J-M Bollag (1997) Isolation and characterization of a chlorinated-pyridinol-degrading bacterium. Appl Environ Microbiol 63: 4096–4098. Fitzgerald JW (1976) Sulfate ester formation and hydrolysis: a potentially important yet often ignored aspect of the sulfur cycle of aerobic soils. Bacteriol Rev 40: 698–721. Fitzgerald JW, HW Maca, FA Rose (1979) Physiological factors regulating tyrosine-sulphate sulphohydrolase activity in Comamonas terrigena: occurrence of constitutive and inducible enzymes. J Gen Microbiol 111: 407–415. French CE, S Nicklin, NC Bruce (1996) Sequence and properties of pentaerythritol tetranitrate reductase from Enterobacter cloacae PB 2. J Bacteriol 178: 6623–6627. Hales SG, GK Watson, KS Dodson, GF White (1986) A comparative study of the biodegradation of the surfactant sodium dodecyltriethoxy sulphate by four detergent-degrading bacteria. J Gen Microbiol 132: 953–961. Higgins TP, JR Snape, GF White (1993) Comparison of pathways for biodegradation of monomethyl sulphate in Agrobacterium and Hyphomicrobium species. J Gen Microbiol 139: 2915–2920. Kahnert A, MA Kertesz (2000) Characterization of a sulfur-regulated oxygenative alkylsulfatase from Pseudomonas putida S-313. J Biol Chem 275: 31661–31667. Kahnert A, P Vermeli, C Wietek, P James, T Leisinger, MA Kertesz (2000) The ssu locus plays a key role in organosulfur metabolism in Pseudomonas putida S-313. J Bacteriol 182: 2869–2878. Kertesz MA (1999) Riding the sulfur cycle—metabolism of sulfonates and sulfate esters by Gram-negative bacteria. FEMS Microbiol Rev 24: 135–175. Li J, J-D Gu, L Pan (2005) Transformation of dimethyl phthalate, dimethyl isophthalate, and dimethylterephthalate by Rhodococcus rubber Sa and modeling the process using the modified Gompertz model. Int Biodet Biodeg 55: 223–232.

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Marshall SJ, GF White (2001) Complete denitration of nitroglycerin by bacteria isolated from a washwater soakaway. Appl Environ Microbiol 67: 2622–2626. Meng M, W-Q Sun, LA Geelhaar, G Kumar, AR Patel, GF Payne, MK Speedie, JR Stacy (1995) Denitration of glycerol trinitrate by resting cells and cell extracts of Bacillus thuringiensis/cereus and Enterobacter agglomerans. Appl Environ Microbiol 61: 2548–2553. Munnecke DM, LM Johnson, HW Talbot, S Barik (1982) Microbial metabolism and enzymology of selected pesticides. In Biodegradation and Detoxification of Environmental Pollutants (Ed AM Chrakrabarty), pp. 1–32. CRC Press, Boca Raton, FL. Pogorevc M, K Faber (2003) Purification and characterization of an inverting stero- and enantioselective sec-alkylsulfatase from the Gram-positive bacterium Rhodoccus ruber DSM 44541. Appl Environ Microbiol 69: 2810–2815. Recksiek M, T Selmer, T Dierks, B Schmidt, K von Figura (1998) Sulfatases, trapping of the sulfonated enzyme intermediate by substituting the active site formylglycine. J Biol Chem 273: 6096–6103. Servent D, C Ducrorq, Y Henry, A Guissani, M Lenfant (1991) Nitroglycerin metabolism by Phanerochaete chrysosporium: evidence for nitric oxide and nitrite formation. Biochim Biophys Acta 1074: 320–325. Singh BK, A Walker, JAW Morgan, DJ Wright (2004) Biodegradation of chloropyrifos by Enterobacter strain B-14 and its use in bioremediation of contaminated soils. Appl Environ Microbiol 70: 4855–4863. Snape JR, NA Walkley, AP Morby, S Nicklin, GF White (1997) Purification, properties, and sequence of glycerol trinitrate reductase from Agrobacterium radiobacter. J Bacteriol 179: 7796–7802. Stuehr DJ, J Santolini, Z-Q Wang, C-C Wei, S Adak (2004) Update on mechanism and catalytic regulation in the NO synthases. J Biol Chem 36167–36170. Sutherland TD, I Horne, RJ Russell, JG Oakeshott (2002) Gene cloning and molecular characterization of a two-enzyme system catalyzing the oxidative detoxification of β-endosulfan. Appl Environ Microbiol 68: 6237–6245. Tehara SK, JD Keasling (2003) Gene cloning, purification, and characterization of a phosphodiesterase from Delftia acidovorans. Appl Environ Microbiol 69: 504–508. Thomas ORT, GF White (1989) Metabolic pathway for the biodegradation of sodium dodecyl sulfate by Pseudomonas spC12B. Biotechnol Appl Biochem 11: 318–327. Wayne LG et al. (1991) Fourth report of the cooperative, open-ended study of slowly growing mycobacteria by the international working group on mycobacterial taxonomy. Int J Syst Bacteriol 41: 463–472. Weir KM, TD Sutherland, I Horne, RJ Russell, JG Oakshott (2006) A singe monooxygenase, Ese, is involved in the metabolism of the organochlorines endosulfan and endosulfate in an Arthrobacter sp. Appl Environ Microbiol 72: 3524–3530. White GF, JR Snape (1993) Microbial cleavage of nitrate esters: defusing the environment. J Gen Microbiol 139: 1947–1957. White GF, JR Snape, S Nicklin (1996) Bacterial biodegradation of glycerol trinitrate. Int Biodet Biodeg 38: 77–82. White GF, KS Dodson, I Davies, PJ Matts, JP Shapleigh, WJ Payne (1987) Bacterial utilisation of short-chain primary alkyl sulphate esters. FEMS Microbiol Lett 40: 173–177. White GF, NJ Russell (1994) Biodegradation of anionic surfactants and related molecules. In Biochemistry of Microbial Degradation (Ed C Ratledge), pp. 143–177. Kluwer Academic Publishers, Dordrecht, The Netherlands.

PART 2 ETHERS AND SULFIDES The degradation of a structurally wide range of ethers under both aerobic and anaerobic conditions is covered in a review (White et al. 1996), which should be consulted for details, particularly of polyethers.

ALIPHATIC AND BENZYLIC ETHERS Aliphatic and benzylic ethers are degraded by hydroxylation of the α-methylene group followed by scission of the ether bond with the formation of an aldehyde and an alkanol (White et al. 1996; Kim and Engesser 2004). In contrast, the degradation of 2-chloroethylvinyl ether by Ancylobacter

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aquaticus is initiated by a dehalogenase, although fission of the C–O–C bond is nonenzymatic (van den Wijngaard et al. 1993). 1. The monooxygenase from Burkholderia cepacia G4/PR1 in which the synthesis of toluene2-monooxygenase is constitutive is able to degrade a number of ethers including diethyl ether and n-butyl methyl ether, though not tert-butyl methyl ether (Hur et al. 1997). 2. tert-Butyl methyl ether (MTBE) has been used as a gasoline additive, and concern has arisen over its biodegradability in view of its water solubility that facilitates dispersion in aquatic systems (Johnson et al. 2000). Although pure cultures have been isolated that are able to mineralize MTBE (Hanson et al. 1999; Hatzinger et al. 2001), there are a number of issues that determine its biodegradation and the pathway that is used: this has been evaluated in dicyclopropylketone-induced and n-octane-grown cells of Pseudomonas putida Gpo1 (Smith and Hyman 2004). Propane-grown cells of Mycobacterium vaccae JOB5 and a strain ENV425 obtained by propane enrichment transformed MTBE to tert-butanol or, by hydroxylation (monooxygenation) of the quaternary methyl group to 2-hydroxyisobutyric acid, which was not, however, used as a growth substrate for the organisms (Figure 11.3) (Steffan et al. 1997). In contrast, Mycobacterium austroafricanum strain IFP 2012 transformed MTBE to tert-butanol, 2-methyl-1,2-propandiol, and 2-hydroxyisobutyrate that was degraded (Ferreira et al. 2006). A pathway for its degradation was proposed that involved a cobalamin-dependent mutase, which converted 2-hydroxyisobutyrate-CoA into 3-hydroxybutyryl-CoA (Rohwerder et al. 2006). 3. Degradations of symmetrical long-chain dialkyl ethers are used to illustrate an entirely different metabolic pathway. The di-n-heptyl-, di-n-octyl-, di-n-nonyl-, and di-n-decyl ethers are degraded by a strain of Acinetobacter sp. to two different groups of metabolites (Figure 11.4): a. To n-heptan-, n-octan-, n-nonan-, and n-decanol-1-acetic acids, which were not metabolized further. b. To glutaric (C5), adipic (C6), pimelic (C7), and suberic (C8) acids, which served as sources of carbon and energy. These compounds were formed by terminal oxidation

CH3 CH3 CH3

FIGURE 11.3

CH3 C

OCH3

CH3

C

OH

+ CH2O

CH3

CH3 CH3

C

OH CO2H

Degradation of methyl tert-butyl ether.

CH3−(CH2)n −O−CH2−CH2−(CH2)n −2 −CH3

CH3−(CH2)n −O−CH2−CH2−(CH2)n −2 −CO2H

CH3−(CH2)n −O−CH2−CO2H

FIGURE 11.4

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HO2C−CH2−(CH2)n −3 −CO2H

Biodegradation of di-n-heptyl ether by Acinetobacter sp.

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Miscellaneous Compounds

4.

5.

6.

7.

8.

9.

575

followed by an unusual oxidation at the carbon atom β to the ether bond (Modrzakowski and Finnerty 1980). A less usual reaction is involved in the degradation of succinyloxyacetate by Zoogloea sp. (Peterson and Llaneza 1974), which is accomplished by a lyase that produced fumarate and glycolic acid. Bis-(1-chloro-2-propyl)ether has two chiral centers and exists in (R,R)-, (S,S)-, and a meso form. It is degraded by Rhodococcus sp. with a preference for the (S,S) enantiomer with the intermediate formation of 1-chloro-propan-2-ol and chloroacetone (Garbe et al. 2006). Although hydrolysis of alkyl sulfates by sulfatases is noted in Part 1 of this chapter, ether cleavage has been shown to be the major pathway for the degradation of dodecyltriethoxy sulfate (Hales et al. 1986). Polyethylene glycol can be metabolized and used for growth by a number of bacteria, and degraded aerobically by sphingomonads including species of Sphingopyxis. Aerobic degradation is generally initiated by oxidation and an unusual flavoprotein dehydrogenase has been characterized from Sphingopyxis (Sphingomonas) terrae (Sugimoto et al. 2001). The mechanism of degradation has not, however, been resolved (White et al. 1996). In contrast, anaerobic degradation involves rearrangement. This has been investigated in a variety of organisms including Pelobacter venetianus (Schink and Stieb 1983), an Acetobacterium sp. (Schramm and Schink 1991), Desulfovibrio desulfuricans, and Bacteroides sp. (Dwyer and Tiedje 1986). The initial product is acetaldehyde, which is formed in two stages by the action of a diol dehydratase and a polyethylene glycol acetaldehyde lyase (Figure 11.5), which is apparently found in all PGE-degrading anaerobic bacteria (Frings et al. 1992). Alkylphenol polyethoxylates, which are widely used as nonionic surfactants, are partially degradable by oxidation and loss of ethoxyethyl groups. Concern has arisen, since, although the number of the ethoxy groups is reduced from about nine or more to two or three (Figure 11.6) (Maki et al. 1994; John and White 1998; Fenner et al. 2002), the metabolites are both appreciably persistent and toxic. Although the alkyl phenols may be formed from the complete oxidation of the polyethoxylate side chains, partially degraded metabolites may apparently be resistant to further degradation (Ball et al. 1989). The degradation of a highly branched nonylphenol polyethoxylate by Pseudomonas putida isolated from activated sludge involved the loss of single ethoxylate groups as acetaldehyde until two ethoxylate residues remained (John and White 1998). The mechanism is reminiscent of that involved in the anaerobic degradation of polyethylene glycols. The metabolism of diethyl ether has been studied in the fungus Graphium sp. strain ATCC 58400, which was able to use this as sole source of carbon and energy. When grown with n-butane, the fungus was able to transform—though not to degrade—tert-butyl methyl HO−(CH2−CH2−O)n −CH2−CH2−OH

HO−(CH2−CH2−O)n −CH(OH)−CH3

HO−(CH2−CH2−O)n-1−CH2−CH2OH + CH3−CHO

FIGURE 11.5 Ar ArO

FIGURE 11.6

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Anaerobic biodegradation of polyethylene glycol. O

(CH2CH2O)n (CH2CH2O)n −1

CH2CH2OH CH2CH2OH

ArO ArO

(CH2CH2O)n

(CH2CH2O)n −1

CH2CO2H CH2CO2H

Partial degradation of alkylphenol polyethoxylates.

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ether to tert-butanol and tert-butyl formate (Hardison et al. 1997). The former is biodegradable under aerobic, and some anaerobic, conditions (Bradley et al. 2002). 10. Rhodococcus sp. strain 219 (Bernhardt and Dickmann 1991) and Pseudonocardia sp. strain K1 (Thiemer et al. 2003) are able to grow at the expense of tetrahydrofuran. It was suggested that degradation was initiated by monooxygenation to 2-hydroxytetrahydrofuran. Mineralization of 1,4-dioxane by Pseudonocardia dioxanivorans strain CB1190 has been described (Parales et al. 1994). Although Pseudonocardia sp. strain ENV4789 that was isolated by enrichment with tetrahydrofuran was unable to grow with 1,4-dioxan, it produced 2-hydroxyethoxyacetate as a metabolite, analogous to that produced from tetrahydrofuran (Vainberg et al. 1994). The fungus Cordyceps sinensis degraded 1,4-dioxane and other cyclic ethers including 1,3-dioxane and tetrahydrofuran. The pathway for 1,4-dioxane involved reductive formation of ethylene glycol, glycolate, and oxalate, although details of the mechanism were not presented (Nakamiya et al. 2005).

ARYL ETHERS Diaryl Ethers Halogenated derivatives of diphenyl ether have been used as herbicides (Scalla et al. 1990) and flame retardants (references in Sellström et al. 1998), and also occur naturally (Voinov et al. 1991). Attention has therefore been directed to this class of compounds that formally includes dibenzofurans and dibenzo[1,4]dioxins, which are discussed in Chapter 10, Part 2. The degradation of diphenyl ether itself by Pseudomonas cepacia has been examined (Pfeifer et al. 1989, 1993) and yields 2-pyrone-6-carboxylate as a stable end product. This may be formed from the initially produced 2,3-dihydroxydiphenyl ether in a reaction formally analogous (Figure 11.7) to that, whereby 3-O-methylgallate is converted into 2-pyrone-4,6-dicarboxylate by 3,4-dihydroxybenzoate 4,5dioxygenase in pseudomonads (Kersten et al. 1982). The degradation of diphenyl ether by a strain of Sphingomonas sp. strain SS33 took place with the fission of both rings (Schmidt et al. 1992), and cells grown with diphenyl ether were able to oxidize dibenzo[1,4]dioxin to 2-(2-hydroxyphenoxy)cis,cis-muconate. After adaptation to growth with 4,4′-difluorodiphenyl ether, the organism grew with the chlorinated, but not the brominated analogs (Schmidt et al. 1993). In general some degree of recalcitrance seems to be associated with halogenated diaryl ethers. Aryl-Alkyl Ethers 1. Phenoxyalkanoates. There has been considerable interest in the persistence of chlorinated phenoxyalkanoates—and particularly of phenoxyacetates and phenoxypropionates, which have been used as herbicides. This has therefore stimulated studies on the degradation of these aryl-alkyl ethers. Considerable effort has been directed to elucidating the subsequent

O

O OH

OH OH

CO2H

+ O

FIGURE 11.7

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O

OH

O

O

CO2H

Biodegradation of diphenyl ether.

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577

steps that culminate in the fission of the aromatic ring and have been discussed in Chapter 9, Part 2. The first step in the degradation of phenoxyalkanoates is dealkylation to the corresponding phenol with the formation of glyoxylate from phenoxyacetates or acetoacetate and acetone from phenoxypropionates. This reaction is mediated by an α-ketoglutaratedependent dioxygenase (Fukumori and Hausinger 1993), and it has been shown that one atom of this is incorporated into pyruvate and succinate in cell extracts of Sphingomonas herbicidovorans by using 18O2 (Nickel et al. 1997). This strain is able to degrade both enantiomers of the racemate (R,S)-4-chloro-2-methylphenoxypropionate (mecoprop) (Zipper et al. 1996), and dioxygenases specific for the R and S enantiomers are induced by growth of cells with the respective enantiomer (Nickel et al. 1997). Enantiospecific dioxygenases involved in the degradation of 2-(2,4-dichlorophenoxy)propionate in Delftia acidovorans have been characterized (Schleinitz et al. 2004). The resulting chlorophenols are then degraded by intradiol fission of 2,4-dichlorocatechol and established pathways (Müller et al. 2004). 2. Aryl methyl ethers. A great deal of attention has been directed to the demethylation of aryl methyl ethers on account of interest in the degradation of lignin and related compounds by both aerobic and anaerobic organisms. a. In aerobic bacteria, the degradation of vanillate, isovanillate, and syringate is initiated by de-O-methylation, and the resulting catechols are then degraded by extradiol dioxygenation (Kasai et al. 2004; Providenti et al. 2006). Three mechanisms for the demethylation of aryl O-methyl ethers have been identified, and are involved in the degradation of methoxylated benzoates. i. The methyl group is converted into CH2O by a dioxygenase vanA and a reductase vanB in Pseudomonas sp. strain HR199 (Priefert et al. 1997) and Acinetobacter sp. (Segura et al. 1999). The sequences of van B have been divided into three clusters (Civolani et al. 2000). ii. A tetrahydrofolate (THF) pathway in which the methyl group is transferred to 5-CH3-THF and then to 5,10-CH2-THF. For example, in Sphingomonas paucimobilis strain SYK-6, the demethylase LigM is required for the degradation of both vanillate and the product of partial O-demethylation of syringate (Abe et al. 2005). In this strain, however, the initial de-O-methylation of 2,2′dihydroxy-3,3′-dimethoxy4,4′-dicarboxybiphenyl is mediated by an unstable oxygenase (Sonoki et al. 2000), so that both mechanisms operate in the same organism. iii. In the degradation of substrates such as syringate (Kasai et al. 2004) and 3,4,5trimethoxybenzoate (Donnelly and Dagley 1980) from which 3-O-methylgallate is formed, the methyl group may be lost as methanol produced by hydrolysis of an intermediate methyl ester. This is discussed in Chapter 7, Part 3. b. Several mechanisms have been elucidated for anaerobic bacteria: i. Different organisms are able to utilize the methyl of the methoxy group for growth, for example, the acetogenic Acetobacterium woodii, although this organism is unable to degrade the aromatic ring (Bache and Pfennig 1981). In Clostridium pfennigii, the methyl group may alternatively be converted into butyrate (Krumholz and Bryant 1985), and Desulfotomaculum thermobenzoicum is able to use methoxylated benzoates in the presence or absence of sulfate (Tasaki et al. 1992). The pathways of de-O-methylation have been elucidated in A. woodii (Berman and Frazer 1992), Sporomusa ovata (Stupperich and Konle 1993), Acetobacterium dehalogenans (Kaufmann et al. 1998), and Moorella thermoacetica, and all of them involve an intermediate THF. The O-demethylase of M. thermoacetica consists of three components: MtvB carries out the O-demethylation, MtvC carries out transfer of the methyl group to Co(III) and MtvA the transmethylation to THF with production

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Environmental Degradation and Transformation of Organic Chemicals HO

HO HO

O

O

HO O

O OMe

OMe

OH

FIGURE 11.8

Biotransformation of a β-aryl ether.

GSH

+

OMe

OMe

O

SG

GSH

OH

HO

OMe

OMe OH

+ GSSG

OH

OMe OH

of CH3-THF that is then metabolized to acetate after incorporation of CO from CO2 (Naidu and Ragsdale 2001). ii. Methanethiol is produced by de-O-methylation in a few organisms. Sporobacterium olearium is able to use a wide range of hydroxylated and methoxylated aromatic compounds for growth (Mechichi et al. 1999). Methanethiol was produced from the methoxy groups and acetate and butyrate from the resulting 3,4,5-trihydroxylated compounds. Sulfide-dependent de-O-methylation has been demonstrated in Parasporobacterium paucivorans that is able to degrade syringate with the production of both methanethiol and dimethyl sulfide, while gallate that is produced from 3,4,5-trimethoxybenzoate is degraded to butyrate and acetate (Lomans et al. 2001). 3. Fission of the aryl–O bond in a β-aryl ether by Sphingomonas paucimobilis SYK-6 involves the operation of four genes: LigD, a dehydrogenase; and three tandem-located glutathione S-transferase genes LigE, LigF, a β-etherase, and LigG, a glutathione lyase (Figure 11.8) (Masai et al. 2003).

SULFIDES, DISULFIDES, AND RELATED COMPOUNDS Interest in the possible persistence of aliphatic sulfides has arisen since they are produced in marine anaerobic sediments, and dimethylsulfide may be implicated in climate alteration (Charlson et al. 1987). Dimethylsulfoniopropionate is produced by marine algae as an osmolyte, and has aroused attention for several reasons. It can be the source of climatically active dimethylsulfide (Yoch 2002), so the role of specific bacteria has been considered in limiting its flux from the ocean and deflecting the products of its transformation into the microbial sulfur cycle (Howard et al. 2006). Sulfides and related compounds may be degraded by a range of different pathways, which are illustrated in the following examples: 1. Bacteria selected for growth with bis-(3-pentfluorophenylpropy)-sulfide as sulfur source are able to use dimethyl sulfoxide, dibenzyl sulfide, and some long-chain disulfides as sources of sulfur (van Hamme et al. 2004). Degradation takes place by oxidation to the sulfone, scission of the C–S bond to an alkanol, and an alkyl sulfinate that is degraded with loss of the sulfur, which is used for growth. 2. The anaerobic reduction of dimethyl sulfide to methane (Oremland et al. 1989), and of dibenzyl disulfide to toluenethiol and finally toluene (Miller 1992) has been described. Methanomethylovorans hollandica is able to grow anaerobically with both dimethyl sulfide and methanethiol (Lomans et al. 1999); whereas dimethyl sulfide, which is used as an electron donor for the growth of Rhodobacter sulfidophilus, is converted into dimethyl sulfoxide (Hanlon et al. 1994).

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Miscellaneous Compounds (a)

579

HS−CH2−CH(NH2)−CO2H

CH2

C

NH2 CO2H

CH3−CO−CO2H S

CH2−CH(NH2)−CO2H

CO2H

S

CH3−C(NH2)−CO2H

NH

CH3

CO2H (b) CH3−S−CH2−CH2−CH(NH2)−CO2H CH2 + (c) (CH3)2S−CH2−CH2−CO2H

CH2

CH3−S−CH2−CH2−CO−CO2H

+ CH3SH + CO2 (CH3)2S + CH2

CH−CO2H

FIGURE 11.9 Elimination reactions during metabolism of aliphatic sulfur compounds: (a) cysteine, (b) methionine, and (c) 2-dimethylsulfoniopropionate.

3. Elimination reactions have been implicated in the degradation of a variety of simple sulfur compounds: a. The desulfhydrase that brings about the degradation of l-cysteine to pyruvate was characterized in Salmonella typhimurium. The low yield of the expected pyruvate could be explained by the formation of 2-methylthiazolidine-2,4-dicarboxylate by reaction of the initially produced 2-aminoacrylate with the substrate (Kredich et al. 1973). d-Cysteine inhibits the growth of Escherichia coli but can be degraded to pyruvate, sulfide, and ammonia by a desulfhydrase that is found in other enteric bacteria including Klebsiella pneumoniae, Enterobacter cloacae, and Citrobacter freundii (Nagasawa et al. 1985). Enzymatic activity involves pyridoxal phosphate, and is able to bring about elimination reactions from other d-cysteine derivatives and from 3-chloro-d-alanine (Nagasawa et al. 1988) (Figure 11.9a). Pyruvate is also formed from l-cysteine by tryptophanase (Newton et al. 1965), and in an analogous reaction from S-(2-aminoethyl)-l-cysteine by Corynebacterium glutamicum (Rossol and Pühler 1992) and Streptococcus anginosus (Yoshida et al. 2002). Exceptionally in a few bacteria, cysteine can apparently be dioxygenated to the sulfinic acid from which pyruvate and sulfite are produced (Dominy et al. 2006). This may putatively be carried out by a lyase that is analogous to that involved in the degradation of the corresponding sulfonate cysteate (Denger et al. 2006). b. The formation of ethene from methionine by Escherichia coli takes place by an elimination reaction (Ince and Knowles 1986) (Figure 11.9b), although considerable complexities have emerged in the control and regulation of this reaction (Mansouri and Bunch 1989). It should be noted that the synthesis of ethene in plants proceeds by an entirely different reaction via S-adenosyl methionine and 1-aminocyclopropane-1carboxylate (Kende 1989). c. Elimination is used for the degradation of dimethylsulfoniopropionate to acrylate and dimethyl sulfide. This has been observed in a strain of Clostridium sp. (Wagner and Stadtman 1962) (Figure 11.9c), and in Alcaligenes sp. strain M3A and Pseudomonas doudoroffii (de Souza and Yoch 1995b). The lyase has been purified from a strain of

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Environmental Degradation and Transformation of Organic Chemicals

an Alcaligenes-like organism (de Souza and Yoch 1995a). The acrylate is metabolized by Alcaligenes sp. strain M3A to β-hydroxypropionate (Ansede et al. 1999, 2001) and by Desulfovibrio acrylicus to propionate. This organism can use both sulfate and acrylate as electron acceptors and a range of electron donors including lactate, succinate, ethanol, propanol, glycerol, glycine, and alanine (van der Maarel et al. 1996a). It is worth noting that whereas this mechanism provides a source of dimethyl sulfide in marine environments, in anaerobic freshwater habitats methyl sulfide (methanethiol) and dimethyl sulfide can be produced by the reaction of sulfide with methoxylated aromatic compounds (Lomans et al. 2001). 4. An alternative pathway for the degradation of dimethylsulfoniopropionate involves successive de-S-methylation to 3-thiomethylpropionate and 3-mercaptopropionate: (CH3)2S+–CH2–CH2–CO2– → CH3–S–CH2–CH2–CO2H → HS–CH2–CH2–CO2H This has been observed in strain BIS-6 that was a versatile methylotroph (Visscher and Taylor 1994) and in strains of Desulfobacterium sp. (van der Maarel et al. 1996b). Exceptionally, both elimination to dimethylsulfide and de-S-methylation to methanethiol have been observed in the roseobacters Silicibacter pomeroyi and Roseovarius nubinhibens. Whereas S. pomeroyi could grow at the expense of dimethylsulfoniopropionate and a number of putative degradation products including acrylate, R. nubinhibens was unable to grow with acrylate and neither strain could utilize 3-mercaptopropionate for growth (González et al. 2003). 5. Dimethyldisulfide is degraded by autotrophic sulfur bacteria with the formation of sulfate and CO2, which then enters the Benson–Calvin cycle (Smith and Kelly 1988). In contrast, dimethyl sulfide and dimethyl sulfoxide are degraded by a strain of Hyphomicrobium sp. by pathways involving the formation from both carbon atoms of formaldehyde, which subsequently enters the serine pathway (Suylen et al. 1986) (Figure 11.10). The key enzyme is methanethiol oxidase, which converts methanethiol into formaldehyde, sulfide, and peroxide (Suylen et al. 1987). A strain of Thiobacillus sp. metabolizes dimethyl sulfide by an alternative pathway involving transfer of the methyl group probably to THF by a cobalamin carrier (Visscher and Taylor 1993). Oxygen is not involved in the removal of the methyl groups, so that the reaction may proceed anaerobically. 6. Strains of some facultatively heterotrophic and methylotrophic bacteria can use CS2 as sole energy source, and under aerobic conditions also COS, dimethyl sulfide, dimethyl disulfide, and thioacetate (Jordan et al. 1995). It was proposed that the strains belonged to the genus Thiobacillus, though they are clearly distinct from previously described species, and they have now been assigned to Paracoccus denitrificans (Jordan et al. 1997). There has been considerable interest in the enantiomeric oxidation of sulfides to sulfoxides and illustrative examples include the following: 1. Enantiomerically pure alkyl aryl sulfoxides have been obtained by the microbial oxidation of the corresponding sulfides (Holland 1988). Both Corynebacterium equi, and fungi

O CH3−S−CH3

CH3−SH CH3−S−CH3 CH2O

FIGURE 11.10

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H−CO2H

CO2

Biodegradation of dimethyl sulfide by Hyphomicrobium sp.

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Miscellaneous Compounds

581

including Aspergillus niger, species of Helminthosporium, and Mortierella isabellina were effective. The same fungi were not, however, able to carry out enantiomeric-selective oxidation of ethylmethylphenyl phosphine apparently due to the intrusion of nonselective chemical autoxidation (Holland et al. 1993). 2. The oxidation of a series of substituted phenylmethyl sulfides was examined in two strains of putatively the same organism—Pseudomonas frederiksbergensis. Two significant features emerged: (a) the enantioselectivity varied widely among the substrates and (b) one of the strains consistently produced sulfoxides with the S configuration, whereas the other produced those with the R configuration (Adam et al. 2005). This is unique among sulfoxide-producing strains.

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Visscher PT, BF Taylor (1994) Demethylation of dimethylsulfoniopropionate to 3-mercaptopropionatre by an aerobic bacterium. Appl Environ Microbiol 60: 4617–4619. Voinov VG, YuN El′kin, TA Kuznetsova, II Mal′tsev, VV Mikhailov, VA Sasunkevich (1991) Use of mass spectrometry for the detection and identification of bromine-containing diphenyl ethers. J Chromatogr 586: 360–362. Wagner C, ER Stadtman (1962) Bacterial fermentation of dimethyl-β-propiothetin. Arch Biochem Biophys 98: 331–336. White GF, NJ Russell, EC Tidswell (1996) Bacterial scission of ether bonds. Microbiol Rev 60: 216–232. Yoch D (2002) Dimethylsulfoniopropionate: its sources, role in the marine food web, and biological degradation to dimethylsulfide. Appl Environ Microbiol 68: 5804–5815. Yoshida Y, Y Nakano, A Amano, M Yoshimura, H Fukamachi, T Oho, Y Koga (2002) lcd from Streptoccus anginosus encodes a C-S lyase with α,β-elimination activity that degrades l-cysteine. Microbiology (UK) 148: 3961–3970. Zipper C, K Nickel, W Angst, H-PE Kohler (1996) Complete microbial degradation of both enantiomers of the chiral herbicide Mecoprop [(R,S)-2-(4-chloro-2-methylphenoxy)]propionic acid in an enantioselective manner by Sphingomonas herbicidovorans sp. nov. Appl Environ Microbiol 62: 4318–4322.

PART 3 ALIPHATIC NITRAMINES AND NITROALKANES NITRAMINES Aliphatic nitramines based on oligomers of –CH2–N–NO2– have been used extensively as explosives. Hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) can be used as a source of nitrogen by several bacteria, including Stenotrophomonas maltophilia strain PB1 (Binks et al. 1995). In Rhodococcus sp. this was accomplished by reduction of the ring, loss of nitrite, and formation of 4-nitro-2,4-diazabutanal as a terminal metabolite that retained one of the nitro groups (Figure 11.11) (Seth-Smith et al. 2002; Fournier et al. 2002). This metabolite is also produced from octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine (HMX) by Phanerochaete chrysosporium (Fournier et al. 2004), although it can be used as a nitrogen source by the facultative methylotroph Methylobacterium sp. strain JS178 (Fournier et al. 2005). RDX can be degraded via methylenedinitramine to CH2O, CH3OH, and CO2 by Klebsiella pneumoniae strain SCZ-1 (Zhao et al. 2002). The use of RDX as a source of both carbon and nitrogen is less common, but this has been observed in Williamsia sp. strain KTR4 and Gordonia sp. strain KTR9 (Thompson et al. 2005). Under anaerobic conditions, the formation of methylenedinitramine is followed by abiotic loss of formaldehyde. Bacterial degradation of the polyazapolycyclic-caged polynitramine (2,4,6,8,10,12-hexanitro-2,4,6,8,10,12-hexaazaisowurtzitane), which will probably displace RDX, has been accomplished when it is used as a source of nitrogen (Trott et al. 2003; Bhushan et al. 2003). Its degradation by salicylate 1-monooxygenase from Pseudomonas sp. strain ATCC 29352 involved successive loss of nitrite followed by hydrolytic reactions and abiotic reactions (Bhushan et al. 2004).

NITROALKANES These contain C–NO2 bonds in contrast to the nitrate esters of, for example, glycerol and pentaerythritol with O–NO2 bonds. Nitroalkanes have been used as solvents, and there are a few naturally occurring nitroalkanes such as the glycoside of the toxic 3-nitropropionic acid and the O-methyl

O2N

N

N

NO2

N

O2N

N

N N

CH2 O2N

NH

NHCHO +

H2N−CHO

NO2

FIGURE 11.11

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H C

CH3 CH3

FIGURE 11.12

C

O

NO2

O NO2

O

CH3 CH3

O H

C

H C

NO2

CH3 CH3

2 CH3COCH3

Metabolism of 2-nitropropane.

ethers of aci-nitro compounds such as enteromycin. It has been suggested that degradation of nitroalkanes may have developed in fungi as a protection against defensive plant toxins. The structurally related aliphatic azoxy compounds are represented by the toxic macrozamin that occurs in plants as a glucoside, the antibiotic elaiomycin, and the microbial metabolite jietacin A that is an insecticide. Only limited attention has been devoted to their biodegradation and is accomplished by different types of oxidation: a. Dioxygenation of 2-nitropropane to acetone has been observed in the yeast Hansenula mrakii (Figure 11.12) (Kido et al. 1978b) and in Neurospora crassa (Gorlatova et al. 1998). Both the neutral and the aci-nitronate may react, and it has been proposed that in N. crassa the reaction proceeds via a flavosemiquinone involving the nitronate radical (Francis et al. 2005). b. A monoxygenase (oxidase) from Fusarium oxysporum oxidized nitroalkanes to the aldehyde with production of H2O2 (Kido et al. 1978a; Gadda et al. 1997), and belongs to the acyl-coenzyme A dehydrogenase superfamily (Daubner et al. 2002). It is also worth drawing attention to the plant metabolites miserotoxin that is a glycoside of nitropropanol, and karakin that is a glucose ester of 3-nitropropionate (Anderson et al. 2005). These exert their toxicity by hydrolysis to nitropropionic acid that is a suicide inactivator of succinate dehydrogenase (Alston et al. 1977). Exceptionally, the aglycones can be used as electron acceptors by Denitrobacterium detoxificans during growing with lactate that is converted into acetate (Anderson et al. 2000). The enantiomeric reduction of 2-nitro-1-phenylprop-1-ene has been studied in a range of Gram-positive organisms including strains of Rhodococcus rhodochrous (Sakai et al. 1985). The enantiomeric purity of the product depended on the strain used, the length of cultivation, and the maintenance of a low pH that is consistent with the later results of Meah and Massey (2000). It has been shown that an NADPH-linked reduction of α,β-unsaturated nitro compounds may also be accomplished by old yellow enzyme via the aci-nitro form (Meah and Massey 2000). This is formally analogous to the reduction and dismutation of cyclic enones by the same enzyme (Vaz et al. 1995), and the reductive fission of nitrate esters by an enzyme homologous to the old yellow enzyme from Saccharomyces cerevisiae (Snape et al. 1997). The transformation of N-nitrosodimethylamine by Pseudomonas mendocina KR1 that has toluene-4-monoxygenase activity was initiated by monooxygenation to the N-nitro compound, which produced N-nitromethylamine and formaldehyde, presumably by hydroxylation of the methyl group (Fournier et al. 2006).

REFERENCES Alston TA, L Mela, HJ Bright (1977) 3-Nitropropionate, the toxic substance of Indigofera, is a suicide inactivator of succinate dehydrogenase. Proc Natl Acad USA 74: 3767–3771.

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Anderson RC, MA Rasmussen, NS Jensen, MJ Allison (2000) Denitrobacterium detoxificans gen. nov., sp. nov., a ruminal bacterium that respires on nitrocompounds. Int J Syst Evol Microbiol 50: 633–638. Anderson RC, W Majak, MA Rasmussen, TR Callaway, RC Beier, DJ Nisbet, MJ Allison (2005) Toxicity and metabolism of the conjugates of 3-nitropropanol and 3-nitropropionioc acid in forage poisonous to livestock. J Agric Food Chem 53: 2344–2350. Bhushan B, A Halasz, J Spain, J Hawari (2004) Initial reaction(s) in biotransformation of CL-20 is catalyzed by salicylate 1-monooxygenase from Pseudomonas sp. strain ATCC 29352. Appl Environ Microbiol 70: 4040–4047. Bhushan B, L Paquet, JC Spain, J Hawari (2003) Biotransformation of 2,4,6,8,10,12-hexanitro-2,4,6,8,10,12hexaazaisowurtzitane (CL-20) by denitrifying Pseudomonas sp. strain FA1. Appl Environ Microbiol 69: 5216–5221. Binks PR, S Nicklin, NC Bruce (1995) Degradation of hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) by Stenotrophomonas maltophilia PB1. Appl Environ Microbiol 61: 1318–1322. Daubner SC, G Gadda, Mp Valley, PF Fitzpatrick (2002) Cloning of nitroalkane oxidase from Fusarium oxysporum identifies a new member of the acyl-CoA dehydrogenase superfamily. Proc Natl Acad Sci USA 99: 2702–2707. Fournier D, A Halasz, J Spain, P Fiurasek, J Hawari (2002) Determination of key intermediates during biodegradation of hexahydro-1,3,5-trinitro-1,3,5-triazine with Rhodococcus sp. strain DN 22. Appl Environ Microbiol 68: 166–172. Fournier D, A Halasz, S Thiboutot, G Ampleman, D Manno, J Hawari (2004) Biodegradation of octahydro1,3,5,7-tetranitro-1,3,5,7-tetrazocine (HMX) by Phanerochaete chrysosporium: new insight into the degradation pathway. Environ Sci Technol 38: 4130–4133. Fournier D, J Hawari, SH Streger, K McClay, PB Hatzinger (2006) Biotransformation of N-nitrosodimethylamine by Pseudomonas mendocina KR1. Appl Environ Microbiol 72: 6693–6698. Fournier D, S Trott, J Hawari, J Spain (2005) Metabolism of the aliphatic nitramine 4-nitro-2,4-diazabutranal by Methylobacterium sp. strain JS 178. Appl Environ Microbiol 71: 4199–4202. Francis K, B Russell, G Gadda (2005) Involvement of a flavosemiquinone in the enzymatic oxidation of nitroalkenes catalyzed by 2-nitropropane dioxygenase. J Biol Chem 280: 5195–5204. Gadda G, RD Edmonson, DH Russell, PF Fitzpatrick (1997) Identification of the naturally occurring flavin of nitroalkane oxidase from Fusarium oxysporum as a 5-nitrobutyl-FAD and conversion of the enzyme to the active FAD-containing form. J Biol Chem 272: 5563–5570. Gorlatova N, M Tchorzewski, T Kurihara, K Soda, N Esaki (1998) Purification, characterization, and mechanism of a flavin mononucleotide-dependent 2-nitropropane dioxygenase from Neurospora crassa. Appl Environ Microbiol 64: 1029–1033. Kido T, K Hashizume, K Soda (1978a) Purification and properties of nitroalkane oxidase from Fusarium oxysporium. J Bacteriol 133: 53–58. Kido T, K Sida, K Asada (1978b) Properties of 2-nitropropane dioxygenase of Hansenula mrakii. J BiolChem 253: 226–232. Meah Y, V Massey (2000) Old yellow enzyme: stepwise reduction of nitroolefins and catalysis of acid-nitro tautomerization. Proc Natl Acad Sci USA 97: 10733–10738. Sakai K, A Nakazawa, K Kondo, H Ohta (1985) Microbial hydrogenation of nitroolefins. Agric Biol Chem 49: 2231–2236. Seth-Smith HMB, SJ Rosser, A Basran, ER Travis, ER Dabbs, S Nicklin, NC Bruce (2002) Cloning, sequencing, and characterization of the hexahydro-1,3,5-trinitro-1,3,5-triazine degradation gene cluster from Rhodococcus rhodochrous. Appl Environ Microbiol 68: 4764–4771. Snape JR, NA Walkley, AP Morby, S Nicklin, GF White (1997) Purification, properties, and sequence of glycerol trinitrate reductase from Agrobacterium radiobacter. J Bacteriol 179: 7796–7802. Thompson KT, FH Crocker, HL Fredrickson (2005) Mineralization of the cyclic nitramine explosive hexahydro-1,3,5-trinitro-1,3,5-triazine by Gordonia and Williamsia spp. Appl Environ Microbiol 71: 8265–8272. Trott S, SF Nishino, J Hawari, JC Spain (2003) Biodegradation of the nitramine explosive CL-20. Appl Environ Microbiol 69: 1871–1874. Vaz ADN, S Chakraborty, V Massey (1995) Old yellow enzyme: aromatization of cyclic enones and the mechanism of a novel dismutation reaction. Biochemistry 34: 4246–4256. Zhao J-S, A Halasz, L Paquet, C Beaulieu, J Hawari (2002) Biodegradation of hexahydro-1,3,5-trinitro-1,3,5triazine and its mononitroso derivative hexahydro-1-nitroso-3,5-dinitro-1,3,5-triazine by Klebsiella pneumoniae strain SCZ-1 isolated from an anaerobic sludge. Appl Environ Microbiol 68: 5336–5341.

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PART 4 ALIPHATIC PHOSPHONATES AND SULFONATES INTRODUCTION Organic phosphonates are both anthropogenic and biogenic. They have been incorporated into household detergents and synthesized as antiviral compounds (acyclic nucleoside phosphonates): they exist in naturally occurring compounds including antibiotics (phosphonomycin [fosfomycin], alaphosphin [alafosfalin], and plumbemycin), in the microbial herbicide bialaphos (phosphinothricin), and in complex cell components. Although alkyl sulfonates have been used as detergents and the aromatic analogs have a number of applications, there are also biogenic alkyl sulfonates including the metabolically important coenzyme M (2-mercaptoethanesulfonate), taurine (2-aminoethanesulfonate) derived from cysteine, and sulfonolipids. It has been shown that sulfonates, which are present in agricultural soil, are important for the survival of Pseudomonas putida in the soil and the rhizosphere (Mirleau et al. 2005). Both phosphonates and sulfonates can be degraded, and the mechanisms of their degradation generally depend on the absence of alternative sources of inorganic phosphorus or sulfur. There are only a few naturally occurring organic compounds containing boron, and these are borate complexes with tetradentate oxygen ligands containing substituted 1,2-dihydroxyethanes. Organic boronates with C–B bonds, particularly aryl boronates are, however, important intermediates in organic synthesis.

PHOSPHONATES Phosphonates can serve as sources of phosphorus, generally in the absence of inorganic phosphate (Cook et al. 1978), although only a single strain may not be able to use all the phosphonates that have been examined (Schowanek and Verstraete 1990). Degradation of phosphonates involves cleavage of the C–P bond with the formation of inorganic phosphate, and may be accomplished by a carbon–phosphorus lyase, although the enzymology and its regulation are extremely complex (Chen et al. 1990). The gene cluster required for the utilization of phosphonates is induced in Escherichia coli by phosphate limitation, and genetic evidence suggests a connection between the metabolism of phosphonates and phosphites. On the basis of this, the interesting suggestion has been made that there may exist a phosphorus redox cycle and that phosphorus is involved not only at the +5 oxidation level, but also at lower oxidation levels (Metcalf and Wanner 1991). Two main pathways are used for the degradation of phosphonates that are initiated by a lyase or a hydrolase (Wanner 1994): 1. C–P lyase has a broad substrate specificity and, for example, dimethyl phosphonate is degraded to methane, methylphenyl phosphonate to benzene, and the degradation of the widely used herbicide glyphosate may follow alternative pathways both of which involve C–P fission. 2. The phosphonatase (hydrolase) pathway is less widely used, typically for the degradation of 2-aminoethylphosphonate via phosphonoacetaldehyde to acetaldehyde, and the degradation of phosphonoacetate that involves a specific hydrolase (Kulakova et al. 2001). Different enzymatic activities have, however, been shown in species of Campylobacter in which these are exceptionally expressed in the presence of high concentrations of phosphate (Mendz et al. 2005). In enteric bacteria and pseudomonads, both or only one of the pathways may be expressed (Jiang et al. 1995), and further examples are given in the illustrations. Examples of the various mechanisms available for the metabolism of phosphonates include the following: 1. In Bacillus cereus, 2-aminoethylphosphonate is initially oxidized to 2-phosphonoacetaldehyde (La Nauze and Rosenberg 1968) before cleavage of the C–P bond (La Nauze et al. 1970) (Figure 11.13). The degradation of 2-aminoethylphosphonate by Pseudomonas

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Miscellaneous Compounds

589 O

H2N−CH2−CH2−P

FIGURE 11.13

O OH

OH

CHO−CH2−P

OH

OH

CH3−CHO + Pi

Biodegradation of 2-aminomethylphosphonate. O CH3 CH3

FIGURE 11.14

OH

O P

P

CH3

OH

OH

+

CH4

Pi

+ CH4

Reductive biodegradation of alkyl phosphonates and phosphites.

O

O (a)

HO2C−CH2−NH−CH2−P

HO2C−CHO +

OH

H2N−CH2−P

OH

OH

OH CH3−NH2 + Pi

O (b)

HO2C−CH2−NH−CH2−P

OH

HO2C−CH2−NH−CH3 + Pi

OH HO2C−CH2−NH2 + CH2O

FIGURE 11.15

Alternative pathways for the biodegradation of glyphosate.

putida strain NG2 is carried out even in the presence of phosphate, and is mediated by pyruvate aminotransferase and phosphonoacetaldehyde hydrolase activities that are induced by 2-aminoethylphosphonate (Ternan and Quinn 1998). In contrast, in Enterobacter aerogenes strain IFO 12010, these activities are induced only under conditions of phosphate limitation. 2. Reductive pathways have been observed in a number of degradations. Klebsiella pneumoniae metabolized methylphenylphosphonate to benzene (Cook et al. 1979). Alkyl phosphonates and phosphites are degraded by a number of bacteria including species of Klyvera and Klebsiella (Wackett et al. 1987a) (Figure 11.14), and by Escherichia coli (Wackett et al. 1987b) using a pathway in which the alkyl groups are reduced to alkanes. 3. In view of its importance as a herbicide, the degradation of glyphosate has been investigated in a number of organisms and two pathways have been elucidated, differing in the stage at which the C–P bond undergoes fission: a. Loss of a C2 fragment—formally glyoxylate—with the formation of aminomethylphosphonate (Pipke and Amrhein 1988), which may be further degraded by cleavage of the C–P bond to methylamine and phosphate (Jacob et al. 1988) (Figure 11.15a). b. Initial cleavage of the C–P bond with the formation of sarcosine, which is then metabolized to glycine (Pipke et al. 1987; Liu et al. 1991) (Figure 11.15b).

SULFONATES Methane sulfonic acid is produced by tropospheric oxidation of methyl sulfides, and there are naturally occurring sulfonates including derivatives of taurine and of glucose-6-sulfonate (sulfoquinovose),

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which are widely distributed in the sulfonolipids of algae and some cyanobacteria. Sulfonates can be degraded by a range of mechanisms, both aerobically and anaerobically (Cook et al. 1999): 1. Strains of methylotrophic bacteria degrade methane sulfonate by initial oxidation to formaldehyde and sulfite (Kelly et al. 1994; Thompson et al. 1995; Baxter et al. 2002). A formally analogous reaction is used in the biodegradation of sulfosuccinate to oxalacetate (Quick et al. 1994). The aerobic strains capable of degrading methanesulfonate have been assigned to the genera Methylosulfonomonas and Marinosulfonomonas (Holmes et al. 1997). Degradation by Methylosulfonomonas methylovora strain M2 is initiated by monooxygenation (hydroxylation) by a multicomponent enzyme. One component is an electron transfer protein of which the larger subunit contains a Rieske [2Fe–2S] center, and both this and the small subunit show a high degree of homology with those of dioxygenase enzymes (De Marco et al. 1999). The degradation of C2 to C10 unsubstituted alkyl sulfonates also takes place by monooxygenation with the formation of sulfite and the corresponding aldehyde (Eichorn et al. 1999). 2. Taurine is degraded aerobically either by a 2-ketoglutarate-dependent dioxygenation to aminoacetaldehyde (Kertesz 1999) (cf. degradation of 2,4-dichlorophenoxyacetate) or by transamination and fission by a lyase that is also used anaerobically with the formation of acetate (Cook et al. 1999). 3. Sulfoacetaldehyde is a central metabolite in the degradation of a number of C2 sulfonates (Cook and Denger 2002), including ethane-1,2-disulfonate by Ralstonia sp. strain EDS1 (Denger and Cook 2001), and taurine by dehydrogenation and transamination (references in Cook and Denger 2002). Degradation of sulfoacetaldehyde is carried out by sulfoacetaldehyde acetyl transferase, also termed sulphoacetaldehyde sulpholyase. The enzyme has been purified and characterized from cells of Alcaligenes defragrans grown with taurine using nitrate as electron acceptor. Its activity required thiamine pyrophosphate, phosphate, and Mg2+ (Ruff et al. 2003), and the reaction involves the C-2 anion of the thiazolium ring and results in the formation of sulfite and acetyl phosphate in a reaction that is formally analogous to the conversion of pyruvate into acetyl-CoA. 4. The aerobic degradation of l-cysteate by Paracoccus pantotrophus is carried out by deamination to 3-sulfolactate from which sulfite is lost with the formation of pyruvate (Cook et al. 2006). The activity of the lyase (3-l-sulfolactatesulfolyase) has been shown to involve pyridoxal 5′-phosphate, and the enzyme has been found in a number of organisms using l-cysteate either as a source of carbon or as an electron acceptor (Denger et al. 2006). 5. a. When lactate is supplied as the carbon source, a few aliphatic sulfonates such as 2-hydroxyethylsulfonate, alanine-3-sulfonate, and acetaldehyde-2-sulfonate are able to serve as sulfur sources and electron acceptors during the anaerobic growth of some sulfate-reducing bacteria (Lie et al. 1996). Several sulfite-reducing species of Desulfitobacterium were able to use 2-hydroxyethanesulfonate as terminal electron acceptor producing acetate and sulfide (Lie et al. 1999). It is relevant to note that species of Desulfitobacterium are also able to use widely different terminal electron acceptors including chlorinated organic compounds, acrylate, and arsenate. b. Anaerobic degradation of alkylsulfonates has been demonstrated in a strain of Bilophila wadsworthia (Desulfovibrio sp.) strain GRZCYSA by different pathways: i. Fermentation of cysteate with the formation of acetate, NH4+, and equimolar amounts of sulfide and sulfate (Laue et al. 1997a). ii. Utilization of aminomethansulfonate and taurine with lactate as electron donor (Laue et al. 1997b). 6. A strain of Rhodopseudomonas palustris, which was isolated by enrichment with taurine, could use this as electron source, and as a source of sulfur and nitrogen during photoautotrophic growth with CO2. Taurine was metabolized to sulfoacetaldehyde and acetyl phosphate by a pathway, which has already been noted (Novak et al. 2004).

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It has already been noted (Chapter 9, Part 4) that the degradation of aromatic sulfonates when they are used as source of carbon involves dioxygenation, whereas when they serve as a source of sulfur in the absence of sulfate, degradation takes place by monooxygenation to produce the corresponding phenol and sulfite (Kertesz 1999).

BORONATES The C–B bond in boronates is fissioned by monooxygenases in both alkyl (Latham and Walsh 1986), and in aryl boronates including substituted phenylboronates and naphthylboronates with production of the corresponding phenols (Negrete-Raymond et al. 2003).

REFERENCES Baxter NJ, J Scanlan, P De Marco, AP Wood, JC Murrell (2002) Duplicate copies of genes encoding methanesulfonate monooxygenase in Marinosulfonomonas methylotrophica strain TR3 and detection of methanesulfonate utilizers in the environment. Appl Environ Microbiol 68: 289–296. Chen C-M, Q-Z Zhuang, Z Zhu, BL Wanner, CT Walsh (1990) Molecuar biology of carbon-phosphorus bond cleavage Cloning and sequencing of the phn (psiD) genes involved in alkylphosphonate uptake and C-P lyase activity in Escherichia coli. J Biol Chem 265: 4461–4471. Cook AM, CG Daughton, M Alexander (1978) Phosphonate utilization by bacteria. J Bacteriol 133: 85–90. Cook AM, CG Daughton, M Alexander (1979) Benzene from bacterial cleavage of the carbon-phosphorus bond of phenylphosphonates. Biochem J 184: 453–455. Cook AM, H Laue, F Junker (1999) Microbial desulfonation FEMS Microbiol Rev 22: 399–419. Cook AM, K Denger (2002) Dissimilation of the C2 sulfonates. Arch Microbiol 179: 1–6. Cook AM, K Denger, THM Smits (2006) Dissimilation of C3-sulfonates. Arch Microbiol 185: 83–90. De Marco P, P Moradas-Ferreira, TP Higgins, I McDonald, EM Kenna, JC Murrell (1999) Molecular analysis of a novel methanesulfonic acid monooxygenase from the methylotroph Methylosulfonomonas methylovora. J Bacteriol 181: 2244–2251. Denger K, AM Cook (2001) Ethanedisulfonate is degraded via sulfoacetaldehyde in Ralstonia sp. strain EDS1. Arch Microbiol 176: 89–95. Denger K, THM Smits, AM Cook (2006) l-Cysteate sulpho-lyase, a widespread pyridoxal 5′-phosphatecoupled desulfonative enzyme purified from Silicibacter pomeroyi DSS-3. Biochem J 394: 657–664. Eichorn E, JR van der Ploeg, T Leisinger (1999) Characterization of a two-component alkanesulfonate monooxygenase from Escherichia coli. J Biol Chem 274: 26639–26646. Holmes AJ, DP Kelly, SC Baker, AS Thompson, P de Marco, EM Kenna, JC Murrell (1997) Methylosulfonomonas methylovora gen. nov., sp. nov., and Marinosulfonomonas methylotropha gen. nov., sp. nov.: novel methylotrophs able to grow on methansulfonic acid. Arch Microbiol 167: 46–53. Jacob GS, JR Garbow, LE Hallas, NM Kimack, GN Kishore, J Schaefer (1988) Metabolism of glyphosate in Pseudomonas sp strain LBr. Appl Environ Microbiol 54: 2953–2958. Jiang W, WW Metcalf, K-S Lee, BL Wanner (1995) Molecular cloning, mapping, and regulation of the pho regulon genes for phosphonate breakdown by the phosphonatase pathway of Salmonella typhimurium LT2. J Bacteriol 177: 6411–6421. Kelly DP, SC Baker, J Trickett, M Davey, JC Murrell (1994) Methanesulphonate utilization by a novel methylotrophic bacterium involves an unusual monooxygenase. Microbiology (UK) 140: 1419–1426. Kertesz MA (1999) Riding the sulfur cycle—metabolism of sulfonates and sulfate esters by Gram-negative bacteria. FEMS Microbiol Rev 24: 135–175. Kulakova AN, LA Kulakov, NY Akulenko, VN Ksenzenko, JTG Hamilton, JP Quinn (2001) Structural and functional analysis of the phosphonoacetate hydrolase (phnA) gene region in Pseudomonas fluorescens 23F. J Bacteriol 183: 3268–3275. La Nauze JM, H Rosenberg (1968) The identification of 2-phosphonoacetaldehyde as an intermediate in the degradation of 2-aminoethylphosphonate by Bacillus cereus. Biochim Biophys Acta 165: 438–447. La Nauze JM, H Rosenberg, DC Shaw (1970) The enzymatic cleavage of the carbon–phosphorus bond: purification and properties of phosphonatase. Biochim Biophys Acta 212: 332–350. Latham J, AC Walsh (1986) Retention of configuration in oxidation of a chiral boronic acid by the flavoenzyme cyclohexanone oxygenase. J Chem Soc Chem Commun 527–528. Laue H, K Denger, AM Cook (1997a) Fermentation of cysteate by a sulfate-reducing bacterium. Arch Microbiol 168: 210–214.

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Laue H, K Denger, AM Cook (1997b) Taurine reduction in anaerobic respiration of Bilophila wadsworthia RZATAU. Appl Environ Microbiol 63: 2016–2021. Lie TJ, T Pitta, ER Leadbetter, W Godchaux, JR Leadbetter (1996) Sulfonates: novel electron acceptors in anaerobic respiration. Arch Microbiol 166: 204–210. Lie TJ, W Godchaux, JR Leadbetter (1999) Sulfonates as terminal electron acceptors for growth of sulfite-reducing bacteria (Desulfitobacterium spp.) and sulfate-reducing bacteria: effects of inhibitors of sulfidogenesis. Appl Environ Microbiol 65: 4611–4617. Liu C-M, PA McLean, CC Sookdeo, FC Cannon (1991) Degradation of the herbicide glyphosate by members of the family Rhizobiaceae. Appl Environ Microbiol 57: 1799–1804. Mendz GL, F Megraud, V Korolik (2005) Phosphonate catabolism by Campylobacter spp. Arch Microbiol 183: 113–120. Metcalf WW, BL Wanner (1991) Involvement of the Escherichia coli phn (psiD) gene cluster in assimilation of phosphorus in the form of phosphonates, phosphite, P i esters, and Pi. J Bacteriol 173: 587–600. Mirleau P, R Wogelius, A Smith, MA Kertesz (2005) Importance of organosulfur utilization for survival of Pseudomonas putida in soil and rhizosphere. Appl Environ Microbiol 71: 6571–6577. Negrete-Raymond AC, B Weder, LP Wackett (2003) Catabolism of arylboronic acids by Arthrobacter nicotinovorans strain PBA. Appl Environ Microbiol 69: 4263–4267. Novak, RT, RF Gritzer, ER Leadbetter, W Godchaux (2004) Phototrophic utilization of taurine by the purple nonsulfur bacteria Rhodopseudomonas palustris and Rhodobacter sphaeroides. Microbiology (UK) 150: 1881–1891. Pipke R, N Amrhein (1988) Degradation of the phosphonate herbicide glyphosate by Arthrobacter atrocyaneus ATCC 13752. Appl Environ Microbiol 54: 1293–1296. Pipke R, N Amrhein, GS Jacob, J Schaefer, GM Kishore (1987) Metabolism of glyphosate in an Arthrobacter sp GLP-1. Eur J Biochem 165: 267–273. Quick A, NJ Russell, SG Hales, GF White (1994) Biodegradation of sulphosuccinate: direct desulphonation of a secondary sulphonate. Microbiology (UK) 140: 2991–2998. Ruff J, K Denger, AM Cook (2003) Sulphoacetaldehyde acetyltransferase yields acetyl phosphate: purification from Alcaligenes defragrans and gene clusters in taurine degradation. Biochem J 369: 275–285. Schowanek D, W Verstraete (1990) Phosphonate utilization by bacterial cultures and enrichments from environmental samples. Appl Environ Microbiol 56: 895–903. Ternan NG, JP Quinn (1998) Phosphate starvation-independent 2-aminoethylphosphonic acid biodegradation in a newly isolated strain of Pseudomonas putida, NG2. System Appl Microbiol 21: 346–352. Thompson AS, NJP Owens, JC Murrell (1995) isolation and characterization of methansulfonic acid-degrading bacteria from the marine environment. Appl Environ Microbiol 61: 2388–2393. Wackett LP, BL Wanner, CP Venditti, CT Walsh (1987b) Involvement of the phosphate regulon and the psiD locus in carbon-phosphorus lyase activity of Escherichi coli K-12. J Bacteriol 169: 1753–1756. Wackett LP, SL Shames, CP Venditti, CT Walsh (1987a) Bacterial carbon-phosphorus lyase: products, rates and regulation of phosphonic and phosphinic acid metabolism. J Bacteriol 169: 710–717. Wanner BL (1994) Molecular genetics of carbon–phosphorus bond cleavage in bacteria. Biodegradation 5: 175–184.

PART 5 DEGRADATION OF ORGANIC COMPOUNDS OF METALS AND METALLOIDS Although knowledge on the biodegradation of these compounds is sparse, a number of them are important in industrial processes. Formation of methylated derivatives may take place in metals and metalloids belonging to groups 15 and 16 of the periodic table, and a few of group 14. These have been discussed in a critical review (Thayer 2002) and in Chapter 3, Part 4, and they have been noted in the context of the bacterial resistance to metals and metalloids. Since carbon monoxide has been considered as an organic compound (Chapter 7, Part 1), it is consistent to make brief comments on metal carbonyls. There are distinct structural types of organic compounds containing metals and metalloids. The first contain covalent carbon–metal bonds and are strictly organometallic compounds, for example, the alkylated compounds of Hg, Sn and Pb, and of Li, Mg, and Al (and formerly Hg), which have been extensively used in laboratory organic synthesis, and Al(C2H5)3 that is a component of the

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Ziegler–Natta catalyst for polymerization of alkenes. Considerable attention has been directed to double-bonded Fischer carbenes of Cr and W, the Schrock carbenes of Ta and Ti, and cyclic polyene ligands of Fe, Co, Cr, and U. Carbonyls of transition metals from groups 6 to 10 of the periodic table include both the monomeric compounds such as Cr(CO)6, Fe(CO)5, Ni(CO)4 and those with two metal groups such as Mn2(CO)10 and Co2(CO)8, which is used industrially for hydroformylation. Although their source has not been identified, it has been shown that volatile compounds from landfills contain carbonyls of Mo and W (Feldmann and Cullen 1997). Metals may also be linked through an oxygen or nitrogen atom to form a stable metal complex without a carbon–metal bond. These include metal complexes of ethylenediamine tetraacetate (EDTA), diethylenetriamine pentaacetate (DTPA), or ethylenediamine tetramethylphosphonate (EDTMP). Metalloid compounds include antimonyl gluconate and bismuth salicylate. Organic compounds of metals and metalloids are both anthropogenic and produced naturally as methylated compounds of which methylmercury has received considerable attention in view of its established toxicity. In addition, resistance to toxic metal cations and metalloid oxyanions by methylation is well established and is discussed in Chapter 3, Part 4. Phenylmercury and organic tin compounds have been used as biocides; methylcyclopentadienyl manganese tricarbonyl has been proposed as a fuel additive. Methylarsonate and dimethylarsinate have been used as herbicides and insecticides, and several organic arsenic compounds have been examined as chemotherapeutic agents. These include phenylarsonic acids and arsenobenzene of which the 3,3′-diamino-4,4′-dihydroxy derivative known as Salvarsan occupied an important place in the history of chemotherapy. A range of organoarsenic compounds including arsenolipids and arsenocarbohydrates that are degraded to arsenobetaine is found in marine biota. A number of the less common metals including Gd, Sm, In, Tc, Au, and the platinum metals have achieved importance in clinical medicine, and include the γ-emitters 99mTc, 111In, and 153Sm with short half-lives (6.0 h, 2.8 d, and 1.93 d, respectively). They are generally administered as organic complexes— the DTPA of 111In in brain imaging and of Gd that is used in MRI; the EDTMP of 153Sm; the cis-diammine-(1,1-cyclobutanedicarboxylato)platinum (carboplatin); and the gold complex of (Et3P)=Au–X, where X is 1-β-thiopenta-O-acetylglucose (auranofin) (Thompson and Orvig 2003). Little is known, however, of the ultimate fate of these metal complexes (Kümmerer and Helmers 2000). Although the degradation or transformation of some organometallic compounds has been examined, only a few of the pathways and the enzymology have been unambiguously established. Only a brief account of the important issues is attempted, but coverage of essentially ecological aspects has not been attempted. Redox systems that are involved in bacterial reactions involving inorganic arsenic have been reviewed (Silver and Phung 2005), and reactions in which selenate and arsenate serve as electron acceptors for growth in the absence of oxygen have been discussed in Chapter 3, Part 2.

TIN The aerobic transformation of tributyltin by bacteria takes place by successive loss of butyl groups to monobutyltin and eventually to inorganic tin (Kawai et al. 1998). A comparable degradation takes place with fungi and yeasts, but is complicated by the simultaneous methylation of the products (Errécalde et al. 1995). Triphenyltin chloride is degraded to diphenyltin and benzene (Inoue et al. 2000), apparently mediated by pyoverdine siderophores produced by the fluorescent pseudomonad (Inoue et al. 2003). The analogous pyochelin has been implicated in the degradation of triphenyltin chloride by Pseudomonas aeruginosa (Sun et al. 2006), and the role of the ferric complex of pyochelin in generating hydroxyl radicals has been demonstrated using EPR (Sun and Zhong 2006).

LEAD It has been shown using (1-14C-ethyl) tetraethyl lead that this was biodegradable, and that the rate of mineralization was adversely affected by the presence of hydrocarbons that are generally simultaneous contaminants (Mulroy and Ou 1998).

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MERCURY In Hg-resistant bacteria that are resistant to organic forms of Hg such as phenylmercuric acetate and methylmercury chloride, lyases are involved in the fission of the C–Hg to form Hg2+ and benzene or methane, and the enzyme has been partly purified (Schottel 1978). The Hg2+ may then be reduced to nontoxic Hg0. The situation under anaerobic conditions for sulfate-reducing bacteria is complicated by the possibility of both methylation and demethylation in the same strain (Pak and Bartha 1998), plausibly by operation of the acetyl-CoA pathway (Choi et al. 1994; Ekstrom et al. 2003). Under anaerobic conditions, demethylation, though not methylation, has been reported for a methanogen (Pak and Bartha 1998).

ARSENIC Tamaki and Frankenberger (1992) quote the degradation of methanearsonic acid to arsenate and CO2 by aerobic bacteria, and growth at the expense of organoarsenic compounds has been reported. Growth with 2-aminoethylarsonic acid has been demonstrated and the products identified as alanine and arsonoacetaldehyde, the latter decomposing spontaneously by hydrolysis (Lacoste et al. 1992). Growth on arsenoacetate has also been reported without, however, identifying the enzyme responsible for scission of the C–As bond (Quinn and McMullen 1995), and strains of Campylobacter degraded both phenylphosphonate and phenylarsonate (Mendz et al. 2005). Dimethylselenide is degraded to CH4 and CO2 by a methylotrophic methanogen grown with dimethylsulfide (Oremland and Zehr 1986), and serves as a source of selenium for Methanocccus voltae during selenium limitation (Niess and Klein 2004).

REFERENCES Choi S-C, TT Chase, R Bartha (1994) Metabolic pathways leading to mercury methylation in Desulfovibrio desulfuricans LS. Appl Environ Microbiol 60: 4072–4077. Ekstrom EB, FM Morel, JM Benoit (2003) Mercury methylation independent of the acetyl-coenzyme A pathway in sulfate-reducing bacteria. Appl Environ Microbiol 69: 5414–5422. Errécalde O, M Astruc, G Maury, R Pinel (1995) Biotransformation of butyltin compounds using pure strains of microorganisms. Appl Organomet Chem 9: 23–28. Feldmann J, WR Cullen (1997) Occurrence of volatile transition metal compounds in landfill has: synthesis of molybdenum and tungsten carbonyls in the environment. Environ Sci Technol 31: 2125–2129. Inoue H, O Takimura, H Fuse, K Murakami, K Kamimura, Y Yamaoka (2000) Degradation of triphenyltin by a fluorescent pseudomonad Appl Environ Microbiol 66: 3492–3498. Inoue H, O Takimura, K Kawaguchi, T Nitoda, H Fuse, K Murakami, Y Yamaoka (2003) Tin-carbon cleavage of organotin compounds by pyoverdine from Pseudomonas chlororaphis. Appl Environ Microbiol 69: 878–883. Kawai S, Y Kurokawa, H Harino, M Fukushima (1998) Degradation of tributyltin by a bacterial strain isolated from polluted river water. Environ Pollut 102: 259–263. Kümmerer K, E Helmers (2000) Hospital effluents as a source of gadolinium in the aquatic environment. Environ Sci Technol 34: 573–577. Lacoste A-M, C Dumora, BRS Ali, E Neuzil, HBF Dixon (1992) Utilization of 2-aminoethylarsonic acid in Pseudomonas aeruginosa. J Gen Microbiol 138: 1283–1287. Mendz GL, F Mégraud, V Korolik (2005) Phosphonate catabolism by Campylobacter spp. Arch Microbiol 183: 113–120. Mulroy, PT, L-T Ou (1998) Degradation of tetraethyllead during the degradation of leaded gasoline hydrocarbons in soil. Environ Toxicol Chem 17: 777–782. Niess UM, A Klein (2004) Dimethylselenide demethylation is an adaptaive response to selenium deprivation in the archaeon Methanococcus voltae. J Bacteriol 186: 3640–3648. Oremland RS, JP Zehr (1986) Formation of methane and carbon dioxide from dimethylselenide in anoxic sediments and by a methanogenic bacterium. Appl Environ Microbiol 52: 1031–1036. Pak K-R, R Bartha (1998) Mercury methylation and demethylation in anoxic lake sediments by strictly anaerobic bacteria. Appl Environ Microbiol 64: 1013–1017.

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Quinn JP, G McMullan (1995) Carbon–arsenic bond cleavage by a newly isolated Gram-negative bacterium strain ASV2. Microbiology (UK) 141: 721–727. Schottel JL (1978) The mercuric and organomercurial detoxifying enzymes from a plasmid-bearing strain of Escherichia coli. J Biol Chem 253: 4341–4349. Silver S, LT Phung (2005) Genes and enzymes in bacterial oxidation and reduction of inorganic arsenic. Appl Environ Microbiol 71: 599–608. Sun G-X, J-J Zhong (2006) Mechanism of augmentation of organotin decomposition by ferripyochelin: formation of hydroxyl radical and organotin-iron ternary complex. Appl Environ Microbiol 72: 7264–7269. Sun G-X, W-Q Zhou, J-J Zhong (2006) Organotin decomposition by pyochelin secreted by Pseudomonas aeruginosa even in an iron-sufficient environment. Appl Environ Microbiol 72: 6411–6413. Tamaki S, WT Frankenberger (1992) Environmental biochemistry of arsenic. Rev Environ Contam Toxicol 124: 79–110. Thayer JS (2002) Biological methylation of less-studied elements. Appl Organometal Chem 16: 677–691. Thompson KH, C Orvig (2003) Boon and bane of metal ions in medicine. Science 300: 936–939.

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Section IV Bioremediation

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Bioremediation: Overview

INTRODUCTION Few of the compounds that are the object of bioremediation programs are readily degraded, and although many have been shown to be degradable in controlled laboratory experiments, the cardinal question is whether they are degraded in the specific environment. Biodegradation implies a significant degree of mineralization, and not merely biotransformation (Chapter 2). The general aspects of bioremediation have been discussed by Thomas and Ward (1989), Morgan and Watkinson (1989a,b), Edgehill (1992), and Alexander (1995a). Discussion of the pathways followed during the biodegradation and biotransformation of a wide range of xenobiotics by both bacteria and fungi has already been given in Chapters 7 through 11. This attempts to provide a general background to details, which are discussed in Chapters 13 and 14.

STRATEGIES Three rather different strategies for bioremediation have been employed: these involve treatment in situ, on-site, or off-site, and each has both advantages and disadvantages. The last two offer a more controlled approach that may be based on scale-up of procedures developed in simulation experiments in microcosms. All of them, however, necessitate a full appreciation of the underlying microbiological issues. It should be emphasized that all procedures will of necessity be site-specific since they must take into account both the physical environment and basic microbiological issues. An illustrative example of a problem that may arise in pump-and-treat systems for treating contaminated aquifers is the reduced permeability at the delivery site, which was due to deposition of colloidal material that was mobilized during the treatment (Wiesner et al. 1996). This was apparently enhanced by gas formation from the hydrogen peroxide used to increase the oxygen concentration in the aquifer. Strategies will also be contaminant-specific, and take into account both the pathways and the regulation mechanisms for degradation of the generally complex range of contaminants. For a given substrate, the induction an d maintenance of the degradative enzymes will depend on the environmental conditions (Guerin and Boyd 1992). The presence or addition of readily degradable substrates may enhance the bacterial population, though not necessarily those required for the degradation of recalcitrant compounds. For example, addition of toluene to an aquifer to facilitate the degradation of trichloroethene by indigenous bacteria could result in the dominance of populations that degrade trichloroethene only poorly. In general, the principles that have emerged from the study of aquatic systems, which have been outlined in previous chapters, can be adapted to the terrestrial environment merely with change of emphasis. For example, the cardinal role of association of xenobiotics with organic components of the soil (Bollag and Loll 1983), the concentration of water (water potential), and the potential role of fungi must be taken into consideration. Two main strategies have been used for bioremediation or bioattenuation (although this may fail to take into account stable metabolites): (a) stimulation of the activity of endogenous organisms by the addition of a substrate (bioaugmentation) or (b) inoculation with the active organisms—often isolated from the same site. This chapter attempts to provide an overview of the application of principles, outlined in previous chapters, to the bioremediation of contaminated terrestrial sites and the associated groundwater. Cardinal microbiological processes will be addressed only briefly, and references should be made to previous chapters for metabolic details. There are several reasons for the heightened concern over the increasing volume of solid waste—both industrial and domestic. On the positive side, however, 599

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although the volume of domestic waste is huge, households will diminish its heterogeneity with the adoption of at-source separation and recycling: 1. Some sites occupy valuable land in urban areas, and it is therefore imperative that its reuse for commercial or domestic building be ensured against current or future health hazards. 2. Although the use of landfills is being phased out, they have been extensively used in the past. a. For several reasons, there is widespread environmental concern over the adverse longterm effects of landfills, which have been used for on-site disposal of industrial waste or sludge after biological treatment. Leachate may be discharged into watercourses, while transformation of the original components in the landfill into others may have taken place, which are equally or even more undesirable. b. The sludge from municipal water treatment facilities may have been sent to landfills, which may therefore contain both nondegraded—and presumably more recalcitrant—substrates including terminal metabolites, which may have been produced after deposition.

CONTAMINANTS The range of compounds in nondomestic landfills is extremely wide, and includes not only industrial waste from the manufacture of chemicals but also in some circumstances substantial volumes of military waste containing explosives and chemical warfare agents. It should also be appreciated that continuous atmospheric precipitation will have occurred, and therefore virtually any established atmospheric contaminant may also be present. Indeed, the atmospheric input into Lake Michigan has been found to exceed that from landfills (Hornbuckle et al. 1995); so important interphase partitions should be taken into account. The range of relevant compounds to which attention should be directed is dependent on the availability and sophistication of analytical procedures both for the original compounds and for their transformation products. An illustrative example is provided by the development of chromatographic procedures that have been applied to carbon black and coal tar, which have revealed the presence of polycyclic aromatic hydrocarbons (PAHs) with up to nine rings (Bemgård et al. 1993). Because many of the compounds discussed here are toxic, there is increasing demand for the remediation of contaminated sites and removal of the offending compounds. In this overview, the concept will be enlarged to include remediation of riverine sediments in the vicinity of industrial production and contamination of groundwater by low-molecular-mass compounds that have penetrated the overlying cover. Emphasis is placed on microbiological aspects, whereas engineering considerations that are of cardinal importance lie beyond the expertise of the authors. An attempt is made in Chapter 14 to bring together and summarize (1) a range of sites that have attracted attention; (2) the contaminants involved; (3) the results of laboratory-model and full-scale experiments; (4) the relevant microbiological aspects including the organisms, the pathways of biodegradation and biotransformation; and (5) determinants of biodegradability, particularly bioavailability and oxygen concentration. A review by Bollag (1992) addresses the important issue of fungus-mediated polymerization reactions of phenols and anilines, which are relevant to their bioavailability and therefore their biodegradability. Biological treatment of wastewater and industrial exhaust gases is excluded. In addition, there is an increasing interest in recovery rather than destruction of valuable components in these discharges. Although this chapter is self-contained, reference should be made to sections of preceding chapters for additional details. Undue prominence may have been given to PAHs and related compounds. There are a number of reasons for this that illustrate clearly the interaction of different—yet collectively important—issues.

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1. History. Interest in sites containing PAHs that antedate the introduction of modern synthetic chemicals. Gas works and the production of tar residues are common in countries that used coal as a source of coal gas for lighting and heating. Many of these sites have now been abandoned since the industry has been replaced by the dominance of electricity and oil. 2. Urban planning. For convenience in the distribution of coal gas, the production sites were often placed close to areas of urban populations. As cities expand, these sites are of increasing interest for new building, and the existence of gasworks residues presents a potentially serious problem. 3. Recalcitrance. It has been demonstrated that although some of the low-molecular-mass PAHs such as naphthalene and phenanthrene (and their alkylated homologs) are biodegradable, those with more than four rings generally present an increasing degree of recalcitrance. 4. Health hazard. It was established in the 1920s and 1930s that many components of coal tar were carcinogenic and that this was due to the presence of PAHs containing four to six aromatic rings. Their existence therefore presents a potential threat to residents, and to office and factory users. Previous chapters have been directed primarily to the aquatic environment. The principles may, with appropriate modification, be extended to terrestrial systems. In practice, there is no distinct boundary between terrestrial and aquatic systems. Both are influenced by the level of the water table and the possibility of leaching from the soil phase. Substantial effort has been directed to a wide range of agrochemicals, and a few of these have already been used as illustration in earlier chapters. Some important general conclusions from these studies have a direct bearing on the subject of this chapter: 1. After successive applications of an agrochemical, a population that has enhanced capability for their degradation may develop. This is negative for the functioning of the chemicals, but would be positive from the view of bioremediation (Racke and Coats 1990). 2. Many organic compounds partition into the organic matrix of soils and sediments as a result of both chemical and microbial reactions, which have been discussed in Chapter 4, and these associated residues may be both less toxic and less amenable to biodegradation than the free compounds. This presents an adverse prognosis for bioremediation. 3. The composition of the original contaminants in sediments may have been altered during deposition on land sites, either by biotic or abiotic transformation. Loss may also have occurred through volatilization of the more volatile components, for example, alkylnaphthalenes in PAHs, or partially dechlorinated PCBs (polychlorinated biphenyls) that have been subjected to drying (Bushart et al. 1998). In addition, bioavailability may be altered by cycles of wetting and drying (White et al. 1998), so that the bioremediation of dredged material may be adversely affected. 4. Mere alteration of the structure of the contaminant by biotransformation may not necessarily be acceptable. For example, although anaerobic dechlorination of PCBs is desirable, microbial reduction of 3,3′-dichlorobenzidine, which is an intermediate in the manufacture of dyes, produced benzidine that is both more toxic and more susceptible to dissemination in the aquatic phase (Nyman et al. 1997). Procedures for the improvement of microorganisms for their applicability to treatment programs have been presented by Timmis et al. (1994); but in this overview, only a parenthetical note is made of procedures for the genetic manipulation to improve the effectiveness of bacteria. The account is devoted to organisms indigenous at contaminated sites, or to inoculation with naturally occurring organisms of established metabolic relevance. In many situations, the waste will have lain for

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considerable lengths of time with the result that a natural flora of degradative microorganisms will have developed through enrichment under natural conditions. This is well established for agrochemicals (Racke and Coats 1990), and it is, therefore, highly desirable to optimize the activity of such organisms in bioremediation programs. There is an enormous literature in this area and only a small fraction of this is cited. Only primary publications have been cited although, as far as possible, references to review articles have been given where extensive literature citations may be found. Attention is drawn to overviews that offer valuable and stimulating general perspectives (Errampalli et al. 1997; Strauss 1997).

SITES A number of contaminated sites have been used for illustration, though the extent to which bioremediation has received practical application is variable. For a number of reasons, emphasis is placed on basic microbiological issues that have emerged from controlled laboratory experiments: a. These illustrate many cardinal issues that should be taken into consideration in scale-up to field exploitation. b. They reveal fundamental issues such as the reasons for using biphenyl to facilitate degradation of PCBs, or of phenol or toluene to induce enzymes for the degradation of trichloroethene. c. Mixtures of a number of compounds are often involved, and among groups such as PAHs, PCBs, and polybrominated diphenyl ethers (PBDEs), there are wide variations in the biodegradability of individual components.

CHEMICAL PROCEDURES For the sake of completeness, attention is drawn to physical procedures for removal of contaminants and for their chemical destruction. A valuable review by Hamby (1996) summarizes chemical and physical treatments of soils and contaminated ground and surface waters. Examples include the following: • Supercritical carbon dioxide extraction of hydrocarbon contaminants (Schleussinger et al. 1996) • Extraction by vapor stripping (Siegrist et al. 1995) • Thermal blanket in situ removal and destruction of PCBs (Iben et al. 1996) • The application of sorbents (Verstraete and Devlieger 1996) • An electrokinetic process combined with in situ chemical degradation (Ho et al. 1999) • The use of subcritical water at 275°C (Lagadec et al. 2000) The application of enzymes has been discussed by Bollag (1992), and is not covered here.

PHYTOREMEDIATION There has been considerable interest in the use of plants for bioremediation and this merits a rather extensive discussion. Plants can play an important role in bioremediation for several reasons: (1) they can transport contaminants from the soil, (2) they can metabolize the contaminants after uptake, or (3) they can produce exudes that support microbial activity for degradation of the contaminants. In addition, bacteria can produce metabolites that counter the effect of toxins produced by fungi, and serve as biocontrol agents that diminish the need for the application of agrochemicals. Plant exudates play an important role in supporting the growth and activity of bacteria that carry out the degradation of contaminants in the rhizosphere and rhizoplane (the external surface of roots

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including adhering particles of soil). Bacteria are also of increasing importance as biocontrol agents against plant pathogens in the roots of plants. The use of higher plants in bioremediation has been discussed (Schnoor et al. 1995; McIntyre and Lewis 1997; Salt et al. 1998). There are, however, important limitations to the use of plants, which must be considered. These include a. The extent and rate at which the contaminants are accumulated, taking into account the loss of volatile components b. The means for disposal of the contaminated plants, for example, by incineration with its accompanying problems c. Metabolism of the contaminants including not only transformation but also the cardinal issue of the extent of degradation In summary, the secondary role of plants in supporting degradative bacteria will probably be the most important factor. Although this account does not attempt to cover municipal or industrial waste-treatment systems, a brief note is included on the use of wetland, in which plants can play a significant role. Wetlands are generally constructed from flow-through or static ponds into which aquatic plants have been introduced. They have mostly been examined for the treatment of wastewater from small communities and their success has often been quantified using traditional parameters. There has, however, been interest in using these for the elimination of contaminants including nonpoint sources of agrochemicals. Aquatic plants have included species of Phragmites and Typha, and the range of contaminants included alkylbenzenesulfonates (Huang et al. 2004), pharmaceuticals such as iboprofen, carbamazapine, and clofibric acid (Matamoros et al. 2005), and the insecticide azinphos-methyl (Schulz et al. 2003). The results should be evaluated carefully since a range of factors can contribute to lowering the concentrations of contaminants including sorption to plants and sediment, photolysis, temperature, and water depth in determining the role of aerobic or anaerobic microbial reactions. The accumulation of p,p′-DDT and trans-nonachlor in fish that were present in a wetland, and the potential for accumulation of other chlorinated contaminants illustrated the adverse potential for further dissemination of the contaminants (Barber et al. 2006). Primary Role of Plants In addition to their direct role in the bioremediation of contaminants, plants can play an important role in their dissemination through consumption by ruminants. The detoxification and metabolism of higher plants has been reviewed by Sandermann (1994), and some illustrative examples include the following: 1. Hybrid poplars (Populus deltoides x nigra) have attracted considerable interest since they are able to transport and metabolize—even though they do not generally degrade—diverse xenobiotics: a. Trichloroethene was metabolized to trichloroethanol and trichloroacetate (Newman et al. 1997, 1999). b. Atrazine was metabolized by reactions involving dealkylation and hydrolytic dechlorination to yield 2-hydroxy-4,6-diamino-1,3,5-triazine (Burken and Schnoor 1997). c. Tissue cultures of poplar metabolized the explosive hexahydro-1,3,5-trinitro-1,3, 5-triazine (RDX) by partial reduction to the 1-nitroso- and 1,3-dintroso derivatives, and in the presence of light these were further metabolized to CH2O, CH3OH, and CO2 (Van Aken et al. 2004). d. The fate of [U-14C]TNT was examined both in soil and hydroponic systems (Thompson et al. 1998). Most of substrate that was taken up was not extractable, and the main

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reaction was association with the roots, while uptake in aged soil was significantly lower than in spiked soils. Partial reduction to 2-amino-4,6-dinitrotoluene, 4-amino-2, 6-dinitrotoluene, and 2,4-diamino-6-nitrotoluene took place as a minor reaction. e. In a long-term study at a site contaminated with creosote, there was a reduction in the area of the plume that coincided with root development into the water table. Although loss of naphthalene, acenaphthylene, and acenaphthene were observed, levels of PAHs with four or more rings persisted. It was concluded that water solubility was the limiting factor (Widdowson et al. 2005). 2. The metabolism of xenobiotics by other plants including aquatic plants has been examined: a. Pentachlorophenol is metabolized by the aquatic plant Eichhornia crassipes to a number of metabolites including di-, tri-, and tetrachlorocatechol; 2,3,5-tri- and tetrachlorohydroquinone; pentachloroanisole; and tetrachloroveratrole (Roy and Hänninen 1994). The phenolic products can be compared with those produced during the photochemical and initial stages in the microbiological metabolism of pentachlorophenol, followed by O-methylation. b. Quite complex transformations can be mediated, and the metabolism of phoxim by plant organs and cell suspension of soybean (Glycine max) are given as examples (Höhl and Barz 1995). c. The uptake and biotransformation of benzene from soil and from the atmosphere have been studied in a number of plants. For example, it was shown that in leaves of spinach (Spinacia oleracea), the label from 14C-benzene was found in muconic, fumaric, succinic, malic, and oxalic acids as well as in specific amino acids. In addition, an enzyme preparation in the presence of NADH or NADPH produced phenol (Ugrekhelidze et al. 1997). d. 2,4,6-Trinitrotoluene was reduced by the aquatic plant Myriophyllum spicatum to aminodinitrotoluenes (Pavlostathis et al. 1998), and in axenic root cultures of Catharanthus roseus the initial metabolites 2-amino-4,6-dinitrotoluene and 4-amino-2,6-dinitro toluene were conjugated—probably with C6 units (Bhadra et al. 1999). There are therefore several important unresolved issues before the phytoremediation of TNT can be exploited. These include accumulation in plant tissues (Hughes et al. 1997) and the phytotoxicity of amino metabolites if they are excreted. e. Reed canary grass (Phalaris arundinacea) was grown in liquid culture and exposed to RDX that was metabolized to the potentially toxic 4-nitro-2,4-diazabutanal (Just and Schnoor 2004). This metabolite is also produced from RDX by strains of Rhodococcus sp., and from the homologous octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine (HMX) by Phanerochaete chrysosporium. 3. Plant cultures have been examined for their ability to transform contaminants. Generally, only limited changes take place, such as hydroxylation and the formation of conjugates with phenolic groups: a. The transformation of pentachlorophenol by cultures of wheat (Triticum aestivum) produced tetrachlorocatechol as its glucuronide (Schäfer and Sandermann 1988). b. Cultures of wheat (Triticum aestivum), tomato (Lycopersicon asculentum), lettuce (Lactuca sativa), and a rose variety carried out rather limited transformations of fluoranthene with plant-specific formation of 1-, 3-, and 8-hydroxy compounds that were conjugated (Kolb and Harms 2000). c. Transformation of PCBs was examined in plant cultures (Wilken et al. 1995). The rates of transformation were generally greater for congeners with lower degrees of substitution, although neither pentachlorobiphenyl (PCB 101) nor hexachlorobiphenyl (PCB 153) was metabolized. The cultures of soybean (Glycine max), white clover (Trifolium repens), and some of the grasses were most effective. A range of conjugated hydroxylated

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products was formed from 2-chlorobiphenyl including 3-chloro-2-hydroxybiphenyl, which involved an NIH shift, while 2,5,2′,5′-tetrachlorobiphenyl (PCB 52) produced four monohydroxylated metabolites (Wilken et al. 1995). d. The transformation of 4-n-nonyl[U-14C]phenol was examined in a range of plant cultures. Metabolism took place in the alkyl side chain with the formation of monoand dihydroxylated products, with a substantial fraction of the label occurring in bound residues including starch, protein, lignin, and hemicellulose (Bokern and Harms 1997). 4. The role of plants has emerged from studies with soil slurries at a site with a cover of Paspalum notatum, Lespedeza cuneata, Solidago sp., and Pinus taeda in which increased degradation of trichloroethene to CO2 took place (Walton and Anderson 1990). In addition, prairie grasses can stimulate the effective remediation of PAH-contaminated soil (Aprill and Sims 1990). 5. Endophytic bacteria can colonize plant tissues without bringing about adverse effects on the plant, in contrast to epiphytes that remain in the external environment. Bacteria with the established potential for degradation have been used to improve the tolerance and degradation of contaminants: a. Cuttings of hybrid poplar (Populus trichocarpa x deltoides) were inoculated with a strain of Burkholderia cepacia VM 1468 carrying genes for the constitutive expression of toluene degradation. Compared with the uninoculated control, plant growth increased both in the absence and in the presence of toluene. Although the introduced strains did not establish themselves in the endophytic bacterial community, gene transfer to other bacteria with the Tol+ phenotype took place in the roots and stems (Taghavi et al. 2005). b. The plasmid pTOM of Burkholderia cepacia G4 was introduced into B. cepacia strain L.S.2.4, which is a natural endophyte of yellow lupin (Lupinus luteus), and the resulting strain VM 1330 was used to inoculate plants of yellow lupin. These plants were tolerant of toluene up to a concentration of 1 g/L in hydroponic cultivation and up to 500 mg/L in nonsterile soil (Barac et al. 2004). Secondary Role of Plants As introduction, general comments address the role of low-molecular-mass products exuded by plants and their role in biodegradation. Bacteria growing at the expense of these compounds can be advantageous both to the plants in their neighborhood (rhizosphere and rhizoplane) and for their possible role in the biodegradation of contaminants in the rhizosphere and neighboring areas. Organic compounds that are exuded from plant roots mediate the colonization of the rhizosphere of many crop plants by fluorescent pseudomonads. Although the structures of these compounds have not always been identified, it has been demonstrated that higher plants excrete a range of phenolic compounds (Fletcher and Hegde 1995; Hegde and Fletcher 1996). These metabolites include naringin, catechin, and myricitin (Donnelly et al. 1994), and although carbohydrates may predominate, malate and succinate are of primary importance. In contrast, amino acids are generally of lesser significance except for chemotaxis that is a requisite for virulence of the wild pathogen Ralstonia solanacearum (Yao and Allen 2006). It has been shown that organic sulfur compounds—putatively sulfonates and sulfate esters—are important for the survival of bacteria in the rhizosphere (Mirleau et al. 2005). In addition, mesophilic archaea belonging to the Crenarchaeota have been characterized in extracts of tomato plants by a combination of enrichment and molecular methods (Simon et al. 2005). There may also be an indirect effect of plants. For example, mineralization of PAHs was examined in microcosms prepared with freshwater sediments and planted with the reed Phragmites australis (Jouanneau et al. 2005). Inoculation with the pyrene-degrading Mycobacterium sp. strain 6PY1 resulted in a greater degree of mineralization than in the absence of plants, and this was

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accompanied by the release of phenanthrene-4,5-dicarboxylic acid, which is an established metabolite from the degradation of pyrene. The positive effect of the plants could be attributed to the increased access of oxygen to the sediment and therefore direct stimulation of bacterial growth. Degradative Enzymes 1. The degradation of 2-chlorobenzoate was supported—presumptively involving enzymatic reaction—by exudates from Dahurian wild rye (Elymus dauricus) (Siciliano et al. 1998). It was also shown that concentrations of 2-chlorobenzoate that had aged in soil for 2 years could be reduced by the indigenous microflora (Siciliano and Germida 1999). Experiments have been carried out on the growth of Dahurian wild rye in the presence of bacterial inoculants that did not adversely affect the mineralization of atrazine under various conditions (Burken and Schnoor 1996). But in the absence of plants of poplar hybrids (Populus deltoides nigra DN34), root exudate resulted in only a slight stimulation of the mineralization of 14 C-labeled atrazine; addition of crushed roots provided a more substantial positive effect. 2. The extradiol catechol fission enzyme from Terrabacter sp. strain DBf63 was introduced into Arabidopsis thaliana using Agrobacterium tumefaciens-mediated transformation, and the haloalkane dehalogenase from Rhodococcus sp. strain m15-3 into Nicotiana tabacum. Seedlings were introduced into hydroponic media and incubated either in darkness or under illumination. It was shown that enzymatic activities from plants expressing the apoplasm-targeted enzymes exceeded those for the cytoplasm-targeted enzymes. This is consistent with transport between the apoplast and the roots, and it was suggested on the basis of these rather preliminary experiments that this methodology could be effective for degrading contaminants in the aquatic environment (Uchida et al. 2005). 3. Mammalian cytochrome P450 2E1 was introduced into tobacco plants that were exposed to trichloroethene in hydroponic medium for 5 d. Trichloroethene epoxide was produced initially, and was rearranged to trichloroacetaldehyde, which was then reduced to trichloroethanol. This was found in samples of leaves, stems, and roots, but was absent in the control plants. Trichloroethanol was subsequently transported to the leaves where it was apparently metabolized (Doty et al. 2000). 4. Pseudomonas fluorescens strain F113 is an important biocontrol strain for sugar beet, and a transposon that contained the genes from Burkholderia sp. strain LB400 for the degradation of biphenyl was inserted into the chromosome (Brazil et al. 1995). The resulting hybrid F113pcb was stable under nonselective conditions, and presented a possible candidate for bioremediation, although the level of expression of the bph genes was low. Improvement was accomplished by cloning the bhp operon under control of the nod system in Sinorhizobium meliloti to produce P. fluorescens strain F113L::1180, which grew faster with biphenyl and was able to transform effectively a range of PCB congeners as great as, or greater than, the original Burkholderia sp. strain LB400 (Villacieros et al. 2005). 5. It is well established that one of the pathways for the aerobic degradation of trichloroethene is monooxygenation by the toluene ortho monooxygenase (Nelson et al. 1987). Transposon integration was used to insert the gene from Burkholderia cepacia strain G4 into the chromosome of Pseudomonas fluorescens strain 2-79. The strain was added to the germinated wheat seedling before planting in soil microcosms (Yee et al. 1998). The average degradation using this strain was 63%, and root colonization was comparable to that of the parent strain lacking the monooxygenase genes. 6. Ralstonia eutropha (Alcaligenes eutrophus) strain NH9 is able to degrade 3-chlorobenzene by the “modified ortho” pathway. The cbnA gene that encodes 3-chlorocatechol-1, 2-dioxygenase was introduced into rice plants (Oryza sativa ssp. japonica) under the control of a virus 35S promoter. 3-Chlorocatechol induced dioxygenase activity in the callus of the plants, and leaf tissues oxidized 3-chlorocatechol with the production of 2-chloromuconate

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(Shimizu et al. 2002). It was therefore shown that it is possible to produce transgenic plants with the capability of degrading chlorinated aromatic compounds, which are degraded with the formation of 3-chlorocatechol. The Role of Plant Exudates in Degradation The interaction of plants and bacteria in the root system plays an important role in the remediation of contaminants, and may involve the use of plant exudates to stimulate the growth of degradative bacteria as well as the penetration of the soil by plant roots (Kuiper et al. 2004): 1. Metabolites that include naringin, catechin, and myricitin (Donnelly et al. 1994) were able to induce the growth and metabolism of bacteria that degrade a range of PCB congeners (Donnelly et al. 1994). It has also been shown that enzymes for the biodegradation of some PCB congeners could be induced by a number of terpenoids (Gilbert and Crowley 1997), which may originate from higher plants. There is therefore a plausible relation between metabolites excreted by higher plants and utilization of them to support bacterial degradation of xenobiotics in the rhizosphere. In addition, some coumarin, flavone, flavanol, and flavanone plant metabolites were able to support the growth of anaerobic bacteria during dechlorination of chlorocatechols (Allard et al. 1992). 2. It has been shown that the fine roots of mulberry (Morus sp.) die at the end of the growing season, and that this coincides with the accumulation of flavones (morusin, morusinol, and kuwanon) that have C5 substituents at C-3 and C-8. These flavones were able to support the growth of Burkholderia sp. strain LB400, which is effective in the degradation of a range of PCB congeners. It was therefore suggested that the growth of fine roots and the production of phenolic metabolites should be taken into account in evaluating remediation in the rhizosphere (Leigh et al. 2002). 3. Pseudomonas putida strain PCL 1444 is an efficient colonizer of the roots of Lolium multiflorum, which produces a highly branched root system, and it was evaluated for its ability to utilize root exudates and the ability to degrade naphthalene (Kuiper et al. 2001). The exudate from seeds, seedlings, and roots contained a range of low-molecular-weight fatty acids including citrate, malate, and succinate and carbohydrates, particularly fructose, glucose, and arabinose, which could support growth of the strain. Although concentrations of propionate were low in the exudate, propionate allowed a high rate of expression of the genes for naphthalene degradation. These results illustrated the importance of root colonization by bacteria that used plant exudates for growth (Kuiper et al. 2002). 4. Bacteria associated with the roots of plants have an important potential for the biodegradation of biogenic methane and carbon monoxide. Examples include methanotrophs associated with aquatic vegetation (King 1994) and the oxidation of CO by bacteria associated with freshwater macrophytes (Rich and King 1998). An important example of considerable environmental significance is the generation of methane in rice paddies where the plant roots provide the carbon source for methanogenesis (Lu et al. 2005). Other Functions of Bacteria in the Rhizosphere Considerable effort has been devoted to the application of bacteria as agents for the biocontrol of plant pathogens. This is motivated by the desire to diminish the application of agrochemicals and, in addition, improve soil quality. Bacteria in the rhizosphere are able to fulfill a number of quite different functions: direct stimulation of plant growth and the production of secondary metabolites by strains of Pseudomonas, including 2,4-diacetylphloroglucinol, phenazine-1-carboxylate, and pyrrolnitrin, which are toxic to seed- and root-rotting plant pathogenic fungi. Valuable overviews discuss in detail the roles of fluorescent pseudomonads (Haas and Défago 2005) and fungi,

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principally species of Trichoderma (Harman et al. 2004), in the control of soil pathogens. Bacteria may have a positive influence on plants by a number of different mechanisms: 1. Association with nonsymbiotic diazotrophs that can contribute to the nitrogen nutrition of the plants: a. Species of Azospirillum can penetrate the roots of cereal crops, and species of Burkholderia, which are established human and plant pathogens, have now been shown to be able to fix nitrogen. Such strains have been isolated from the rhizosphere, rhizoplane, and the roots of field crops of maize and sugar in Mexico and Brazil (Perin et al. 2006). b. Stimulating the growth of plants by the production, for example, of auxins by Pseudomonas putida (Patten and Glick 2002) or of quinolinate phosphoribosyl transferase by Burkholderia sp. strain PsJN (Wang et al. 2006). 2. For effective function as biological control agents, there are a number of determinants that must be fulfilled: a. Bacteria must be present on the roots of the plants, be able to produce microcolonies, and synthesize inhibitory compounds. Examples include the production against root infection by fungal pathogens of (i) the antimicrobial 2,4-diacetylphloroglucinol by Pseudomonas fluorescens CHA0 (Bottiglieri and Keel 2006), (ii) pyoluteorin by Pseudomonas fluorescens Pf-5 (Brodhagen et al. 2004), and (iii) phenazine-1-carboxamide by Pseudomonas chlororaphis PCL 1391 (Chin-A-Woeng et al. 2000). The combination of Pseudomonas fluorescens CHA0 and P. fluorescens Q2-87, each of which produces 2,4-diacetylphloroglucinol, resulted in reciprocal stimulation of the expression of 2,4-diacetylphloroglucinol (Maurhofer et al. 2004). Conversely, Fusarium oxysporum can induce resistance to 2,4-diacetylphloroglucinol by deacylation to the less-toxic 2-acetylphloroglucinol and phloroglucinol (Schouten et al. 2004). b. The production of phenazine-1-carboxamide by Pseudomonas chlororaphis PCL 1391 is a requirement for its function in controlling tomato rot caused by Fusarium oxysporum, and quorum sensing mediated by N-hexanoyl-l-homoserine lactone in the medium (Chin-A-Woeng et al. 2001) regulates its synthesis. Its production in this strain is promoted by a number of conditions including the presence of amino acids, especially phenylalanine and tryptophan (Tjeerd van Rij et al. 2004). This is consistent with the biosynthesis of aromatic amino acids and phenazines in branches of the chorismate pathway, whereas both the N-9 and the N-10 heteroatoms of phenazine originate from glutamine (Mavrodi et al. 1998, 2001). c. Degradation of fungal toxins may contribute to the protection of the plants. For example, fusaric acid produced by Fusarium oxysporum can be degraded by Burkholderia spp. (Utsumi et al. 1991). In addition, fusaric acid produced by the fungus may have other adverse effects: i. In the production of phenazine-1-carboxamide, which is required for the suppression of tomato rot by Pseudomonas chlororaphis PCL1391 (Tjeerd van Rij et al. 2005) ii. In the biosynthesis of 2,4-diacetylphloroglucinol in Pseudomonas fluorescens CHA0, which is important for biocontrol in the wheat rhizosphere (Notz et al. 2002) d. It has been shown in Pseudomonas fluorescens CHA0, which is an antagonist of the plant pathogenic nematode Meloidogyne incognita, that the extracellular protease encoded by the gene aprA was involved either directly or indirectly (Siddiqui et al. 2005). e. Phase variation regulates the production of flagellae, surface lipoproteins, secondary metabolites, and hydrolytic enzymes including proteases, and is important in the colonization of the rhizosphere by biocontrol agents as a result of increased motility. This has been examined in Pseudomonas fluorescens strain F113 during the colonization of alfalfa (Medicago sativa). Colony variation was implicated in its effectiveness

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(Sánchez-Contreras et al. 2002), and is mediated by the site-specific recombinases Sss and XerD (Martínez-Granero et al. 2005, 2006). An analogous phenomenon occurred in Pseudomonas sp. strain PCL1171 in which mutations in the genes gacA and gacS were shown to be responsible (van den Broek et al. 2005). 3. Several studies have been directed to Pseudomonas putida strain KT2440, which has attracted interest both as a biocontrol agent and for the degradation of contaminants. It was able to maintain high cell densities when it was introduced as a coating of seeds during growth of agricultural crops (Molina et al. 2000). This organism is able to utilize a number of amino acids including proline, lysine, and glutamate for growth, and attention has been directed to their metabolism. The enzymes in the davDt operon for proline degradation are induced by corn-root exudates (Vílchez et al. 2000). There are several pathways for the degradation of lysine (Revelles et al. 2005), and the operon that is involved in its metabolism is induced by δ-aminovalerate, which is one of the early degradation products (Revelles et al. 2004). Particular attention has been directed to 1-aminocyclopropane-1carboxylate (ACC), which is a key intermediate in the biosynthesis of ethene by plants, and to its role in promoting the growth of agricultural plants: a. It has been shown that Pseudomonas putida strain GR12-2 exhibits ACC deaminase activity, and that this organism promoted the root development of canola (Brassica campestris) (Glick et al. 1997). The specific role of ACC was inferred from the fact that mutants unable to utilize ACC and lacking the deaminase did not exhibit this stimulation (Glick et al. 1994). b. ACC deaminase has been demonstrated in a range of bacteria isolated from the rhizosphere of Pisum sativum and Brassica juncea (Belimov et al. 2001). It has been proposed on the basis of several lines of evidence that the stimulation by ACC deaminase is accomplished in several steps initiated by extrusion of ACC from the plant, which is then removed through bacterial degradation mediated by ACC deaminase. This results in decreased ethene production by the plant and thereby, limitation of growth inhibition. Bacteria in the rhizosphere are therefore able to play an important role in seed and plant development. The Role of Mycorrhizal Fungi Mycorrhizal fungi play an important role in the nutrition and survival of higher plants, and the importance of their associated bacteria has been recognized. In a study of the bacterial communities in the mycorrhizospheres of Pinus sylvestris—Suillus bovinus and P. sylvestris— and Paxillus involutis, the occurrence of both Gram-negative and Gram-positive bacteria was recognized (Timonen et al. 1998). A wide range of taxa were found, including Burkholderia cepacia and Paenibacillus sp., in which strains with established biodegradative capability have been recognized. The relevance of rhizosphere bacteria was illustrated by the demonstration of a biofilm community at the interface of the petroleum-contaminated soil/mycorrhizosphere. Strains of Pseudomonas fluorescens biovars were isolated and these were able to grow at the expense of 1,3-dimethyl benzene and 3-methylbenzoate, brought about fission of catechol, and harbored the plasmid-borne genes xylE and xylM (Sarand et al. 1998). Important conclusions can be drawn collectively from the results of these studies: • Microbial degradation in the root system of higher plants can be significant. • Higher plants play a cardinal role in the provision of organic carbon and of exudates. • The effectiveness of bacteria depends on the nature of the soil and, in particular, its organic content. As a dividend, bacteria are important as biocontrol agents against a number of plant pathogens.

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MICROBIOLOGICAL ASPECTS Essential aspects of the metabolism of xenobiotics are given in earlier chapters. Therefore, only a cursory discussion is given for classes of compounds that have already been reviewed; for others, a few additional comments have been given. Although a valuable review on microbiological aspects of bioremediation of sites contaminated with organochlorine compounds has been given (Morgan and Watkinson 1989b), it is convenient to provide a short summary of some cardinal issues that are equally significant for all xenobiotics. A number of issues that have already been noted in Chapter 4 are discussed here in the context of bioremediation. In natural systems, many different microorganisms will invariably be present, and the degradation of a xenobiotic is frequently dependent upon the activity of several microorganisms. The increased appreciation of the role of populations is considered in Chapter 13.

REGULATION OF PATHWAYS It is very seldom that only a single substrate is present. It is therefore important to examine how the regulation of degradative pathways may be affected and, in particular, whether the simultaneous presence of other contaminants has an adverse effect. In addition, some of the components of a contaminant may directly inhibit degradation by toxification of the relevant organism. The example of azaarenes in groundwater at a wood preservation site that inhibit PAH degradation (Lantz et al. 1997) is noted in Chapter 14. The regulatory pathways for monocyclic aromatic compounds such as phenol, benzoate, and hydroxybenzoates have been reviewed in detail (Ornston and Yeh 1982). Whereas biodegradation of a substrate is frequently induced by growth with an early metabolite, for example, of biphenyl (Furukawa et al. 1983) and naphthalene (Barnsley 1976) by salicylate, the metabolites in the lower pathway may suppress induction of essential degradative enzymes. For example, although the enzymes for the degradation of toluene in Pseudomonas putida strain pWWO are induced during succinate-limited growth, they are suppressed when succinate is in excess (Austen and Dunn 1980; Duetz et al. 1994). Where plausibly comparable degradation pathways exist, however, for example, in the biodegradation of PAHs, a single organism may display wide versatility. For example, Sphingomonas (Pseudomonas) paucimobilis strain EPA505 is capable of degrading fluoranthene, pyrene, benz[a]anthracene, chrysene, benzo[a]pyrene, and benzo[b]fluoranthene (Ye et al. 1996). Degradation of all of them involves the sequential operation of a dioxygenase, a dehydrogenase, and ring-fission enzymes. For halogenated substrates, the situation is more complex. Whereas aerobic degradation of PCBs is often carried out by organisms enriched by growth with biphenyl itself (Hernandez et al. 1995), naphthalene is a much less effective inducer (Pellizari et al. 1996). The degradation of toluene and chlorobenzene generally proceeds by respective extradiol (2:3) and intradiol (1:2) fission of the initially formed catechols. However, since 3-chlorocatechol inhibits the activity of the 2,3-dioxygenase, the degradation of toluene and chlorobenzene are generally incompatible except in strains, such as Pseudomonas sp. strain JS150, which have developed a strategy for overcoming this limitation (Haigler et al. 1992). Further discussion of this issue has been given in Chapter 9, Part 1. A strain of Comamonas testosteroni was, however, able to degrade both 4-chlorophenol and 4-methylphenol (Hollender et al. 1994), and Phanerochaete chrysosporium to degrade simultaneously chlorobenzene and toluene (Yadav et al. 1995). Comparable incompatibility due to the synthesis of toxic metabolites may also be displayed in the degradation of chlorobenzoates (Reineke et al. 1982). An additional example is the synthesis of protoanemonin by organisms that degrade 4-chlorobiphenyl by intradiol fission of 4-chloroatechol (Blasco et al. 1995); protoanemonin may then inhibit the growth of other PCB-degrading organisms in terrestrial systems (Blasco et al. 1997).

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ALTERNATIVE ELECTRON ACCEPTORS In the absence of molecular oxygen, a number of alternative electron acceptors may be used; these include nitrate, sulfate, selenate, carbonate, chlorate, Fe(III), Cr(VI), and U(VI), and have already been discussed in Chapter 3, Part 2. In Chapter 14, which deals with applications, attention is directed primarily to the role of nitrate, sulfate, and Fe(III)—with only parenthetical remarks on Cr(VI) and U(VI). The role of nitrate and sulfate as electron acceptors for the degradation of monocyclic aromatic compounds is discussed, and the particularly broad metabolic versatility of sulfate-reducing bacteria is worthy of notice.

AGING This is a cardinal issue (Hatzinger and Alexander 1995), and the term is applied to processes—whose mechanisms are only incompletely established—whereby organic compounds become associated with polymeric components of the soil matrix with the result that these initially monomeric compounds are no longer accessible to the relevant degradative microorganisms. This implies that evaluations of biodegradability that do not take this into account may seriously underestimate the persistence of the compound. A critical issue is the degree of reversibility of these complexes, since slow release by leaching could introduce the xenobiotic into the aqueous phase and thereby contaminate groundwater and present an additional problem. Although this process has been suggested as a mechanism whereby in the short term the adverse effect of contaminants may be diminished, it seems unwise to conclude that this is an acceptable long-term solution to diminishing their environmental impact (Alexander 1995b).

BIOFILMS These are ubiquitous and only a few brief comments are given here in the context of biodegradation. Both in aquatic and terrestrial systems, bacteria may be present in the form of biofilms that are formed from polymers produced by the organisms themselves. These may both increase their penetration of the matrix and localize the organisms, thereby preventing their loss from the system (Wolfaardt et al. 1995). Their role in wastewater trickling filter treatment systems is, for example, well established, and examples of their possible significance in the biorestoration of contaminated groundwater and a contaminated stream are given in Chapter 14.

METABOLITES AND ENZYMES Metabolite-related parameters have also been used directly to establish biodegradation and biotransformation, and include the following examples: 1. The presence of metabolites determined from laboratory experiments of degradation pathways. Examples include (a) cis-dihydrodiols of PAHs in a marine sediment (Li et al. 1996) and naphthalene in leachate from a contaminated site (Wilson and Madsen 1996), (b) chlorobenzoates from the degradation of PCBs (Flanagan and May 1993), and (c) benzylsuccinates from the anaerobic degradation of benzene, toluene, ethylbenzene, and xylenes (BTEX) (Beller et al. 1995; Beller 2002). 2. Some enzymes have a low substrate specificity, and are able to accept analog substrates with the formation of colored metabolites. This may be used to detect enzymatic activity toward the desired substrates when a large number of colonies have to be examined. Examples include the following: a. For PAHs, toluene 2,3-dioxygenase (Eaton and Chapman 1995) and naphthalene dioxygenase (Ensley et al. 1983) have been used, for which indigo is produced from indole carboxylates.

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b. Toluene-2-monooxygenase for trichloroethene cometabolism by formation of the epoxide (Fries et al. 1998a,b). c. For the degradation of 2,4-dichlorophenoxyacetate, 2,4-dichlorophenoxyacetate/ 2-ketoglutarate dioxygenase, which can accept 4-nitrophenoxyacetate as a substrate with the formation of colored 4-nitrophenol, has been used (Sassenella et al. 1997). Care ought to be exercised, however, since it was also shown that not all strains with established degradative capability for 2,4-dichlorophenoxyacetate were able to form 4-nitrophenol from the surrogate substrate (Sassenella et al. 1997). 3. Examination of a collection of strains of Sphingomonas sp. from a range of localities has suggested that the glutathione S-transferase-encoding gene might be used as a marker for PAH-degrading bacteria (Lloyd-Jones and Lau 1997). In addition, the formation of terminal metabolites may be adverse either for the organism itself, or for other organisms in the ecosystem. Microbial metabolites may also undergo purely chemical reactions to compounds that are terminal products. Examples include the formation of 5-hydroxyquinoline-2-carboxylate from 5-aminonaphthalene-2-sulfonate (Nörtemann et al. 1993) or benzo[b]naphtho[1,2-d]thiophene from benzothiophene (Kropp et al. 1994). Microbial metabolites may be toxic to both the bacteria producing them and to higher organisms. Illustrative examples of toxicity include the following: a. Protoanemonin synthesized during the biodegradation of 4-chlorobiphenyl (Blasco et al. 1997) to PCB-degrading bacteria b. 3,4,3′,4′-Tetrachloroazobenzene produced from 3,4-dichloroaniline (Lee and Kyung 1995) c. The greater carcinogenicity of benzidine produced by anaerobic dechlorination of 3,3′-dichlorobenzidine (Nyman et al. 1997), which is a reduction product of azo dyes and pigments formerly used Metabolites may also play a role in the association of the substrate with humic and fulvic acid components. Two illustrations are given: (a) naphth-1-ol, an established fungal metabolite of naphthalene, may play a role in the association of naphthalene with humic material (Burgos et al. 1996) and (b) it has been shown that 13C-labeled metabolites of [9-13C]-anthracene including 2-hydroxyanthracene-3-carboxylate and phthalate were not extractable from soil with acetone or dichloromethane, and required alkaline hydrolysis for their recovery (Richnow et al. 1998). The identification of volatile compounds of both Mo and W in gases from a municipal landfill (Feldmann and Cullen 1997) illustrated unusual transformation products. Although these compounds were tentatively identified as Mn2(CO)10 and W(CO)6, neither the mechanism of their formation nor their potential health hazards has been resolved.

UTILIZATION OF NITROGEN, SULFUR, AND PHOSPHORUS Many contaminants contain nitrogen, phosphorus, or sulfur, and degradative organisms may utilize one or more of these leaving the major part of the substrate intact. This is particularly important for munitions-related compounds with a high N/C ratio, when the addition of carbon sources may lead to the favorable development of anaerobic or facultatively anaerobic microorganisms.

SUBSTRATE CONCENTRATION There is substantial evidence from the aquatic environment for the existence of threshold concentrations below which rates of degradation of xenobiotics are slow or even negligible (Alexander 1985). A number of hypotheses have been put forward including the critical concentrations required for

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induction and maintenance of the degradative enzymes (Janke 1987). Whether this occurs in the terrestrial environment as well has been explored only superficially, but may be important when only low concentrations of the substrate are available, for example, by desorption. However, organisms that are adapted to such concentrations may be present, and examples have been provided in Chapter 4. Studies with Burkholderia sp. strain PS14 are instructive: 1,2,4,5-tetra- and 1,2,4-trichlorobenzenes at concentrations of 500 nM were mineralized to the extent of 70% in mineral liquid medium within 1 h, and in spiked sterilized soil to ca. 80% after 100 min (Rapp and Timmis 1999). Although in this experiment there was no evidence for a threshold concentration >0.5 nM, the situation in aged soils in which mass transfer is limiting may be significantly less encouraging.

TEMPERATURE Many contaminated sites are located in areas with temperatures considerably less than 15°C, and investigations into the role of temperature have therefore been carried out. Some of the cardinal results have been discussed in Chapter 4, and it may be concluded that although the rates at 4–5°C may be low, selection of organisms for adaptation to the ambient temperature will generally ensure that degradative activity is retained. There may, however, be selective degradation of specific groups of components in mixtures such as PCBs.

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Schleussinger A, B Ohlmeier, I Reiss, S Schulz (1996) Moisture effects on the cleanup of PAH-contaminated soil with dense carbon dioxide. Environ Sci Technol 30: 3199–3204. Schnoor JL, LA Licht, SV McCutcheon, NL Wolfe, LH Carreiora (1995) Phytoremediation of organic and nutrient contaminants. Environ Sci Technol 29: 318A–323A. Schouten A, G van den Berg, C Edel-Hermann, C Steinberg, N Gautheron, C Alabouvette, CH de Vos, P Lemanceau, JM Raaijmakers (2004) Defense responses of Fusarium oxysporum to 2,4-diacetylphloroglucinol, a broad-spectrum antibiotic produced by Pseudomonas fluorescens. Mol Plant–Microbe Interact 17: 1201–1211. Schulz R, C Hahn, ER Bennett, JD Dabrowski, G Thiere, SKC Peall (2003) Fate and effects of azinophosmethyl in a flow-through wetlands in South Africa. Environ Sci Technol 37: 2139–2144. Shimizu M, T Kimura, T Koyama, K Suzuki, N Ogawa, K Miyashita, K Dakka, K Ohmiya (2002) Molecular breeding of transgenic rice plants expressing a bacterial chlorocatechol dioxygenase gene. Appl Environ Microbiol 68: 4061–4066. Siciliano SD, H Holdie, JJ Germida (1998) Enzymatic activity in root exudates of Dahurian wild rye (Elymus daurica) that degrades 2-chlorobenzoic acid. J Agric Food Chem 46: 5–7. Siciliano SD, JJ Germida (1999) Enhanced phytoremediation of chlorobenzoates in rhizosphere soil. Soil Biol Biochem 31: 299–305. Siddiqui IA, D Haas, S Heb (2005) Extracellular protease of Pseudomonas fluorescens CHA0, a biocontrol factor with activity against the root-knot nematode Meloidogyne incognita. Appl Environ Microbiol 71: 5646–5649. Siegrist RL, OR West, MI Morris, DA Pickering, DW Greene, CA Muhr, DD Davenport, JS Gierke (1995) In situ mixed region vapor stripping in low-permeability media. 2. Full-scale field experiments. Environ Sci Technol 29: 2198–2207. Simon HM, CE Jahn, LT Bergerud, MK Sliwinski, PJ Weimer, DK Willis, RM Goodman (2005) Cultivation of mesophilic soil crenarchaeotes in enrichment cultures from plant roots. Appl Environ Microbiol 71: 4751–4760. Strauss HS (1997) Is bioremediation a green technology? J Soil Contam 6: 219–225. Sturchio NC, JL Clausen, IJ Heraty, L Huang, BD Holt, TA Abrajano (1998) Chlorine isotope investigation of natural attenuation of trichloroethene in an aerobic aquifer. Environ Sci Technol 32: 3037–3042. Taghavi S, T Barac, B Greenberg, B Borremans, J Vangronsveld, D van der Lelie (2005) Horizontal gene transfer to endogenous endophytic bacteria from poplar trees improves phytoremediation of toluene. Appl Environ Microbiol 71: 8500–8505. Thomas JM, CH Ward (1989) In situ biorestoration of organic contaminants in the subsurface. Environ Sci Technol 23: 760–766. Thompson PL, LA Ramer, JL Schnoor (1998) Uptake and transformation of TNT by hybrid poplar trees. Environ Sci Technol 32: 975–980. Timmis KN, RJ Steffan, R Unterman (1994) Designing microorganisms for the treatment of toxic wastes. Annu Rev Microbiol 48: 525–557. Timonen S, KS Jörgensen, K Haahtela, R Sen (1998) Bacterial community structure at defined locations of Pinus sylvestris–Suillus bovinus and Pinus sylvestris–Paxillus involutus mycorrhizospheres in dry pine forest humus and nursery peat. Can J Microbiol 44: 499–513. Tjeerd van Rij E, G Girard, BJJ Lugtenberg, GV Bloemberg (2005) Influence of fusaric acid on phenazine-1carboxamide synthesis and gene expression of Pseudomonas chlororaphis strain PCL1391. Microbiology (UK) 151: 2805–2814. Tjeerd van Rij E, M Wesselink, TFC Chin-A-Woeng, GV Bloemberg, BJJ Lugtenberg (2004) Influence of environmental conditions on the production of phenazine-1-carboxamide by Pseudomonas chlororaphis PCL 1391. Mol Plant-Microbe Interact 17: 557–566. Uchida E, T Ouchi, Y Suzuki, T Yoshida, H Habe, I Yamaguchi, T Omori, H Nojiri (2005) Secretion of bacterial xenobiotic-degrading enzymes from transgenic plants by an apoplastic expression system: applicability for phytoremediation. Environ Sci Technol 39: 7671–7677. Ugrekhelidze D, F Korte, G Kvesitadze (1997) Uptake and transformation of benzene and toluene by plant leaves. Ecotoxicol Environ Saf 37: 24–29. Utsumi R et al. (1991) Molecular cloning and characterization of the fusaric acid-resistance gene from Pseudomonas cepacia. Agric Biol Chem 55: 1913–1918. Van Aken B, JM Yoon, CL Just, JL Schnoor (2004) Metabolism and mineralization of hexahydro-1,3, 5-trinitro-1,3,5-triazine inside poplar tissues (Populus deltoides x nigra DN-34). Environ Sci Technol 38: 4572–4579.

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van den Broek D, TFC Chin-A-Wong, GV Bloemberg, BJJ Lugtenberg (2005) Molecular nature of spontaneous modifications in gacS which cause colony phase variation in Pseudomonas sp. strain PCL1171. J Bacteriol 187: 593–600. Verstraete W, W Devlieger (1996) Formation of non-bioavailable organic residues in soil: perspectives for site remediation. Biodegradation 7: 471–485. Vílchez S, L Moloina, C Ramos, JL Ramos (2000) Proline catabolism by Pseudomonas putida: cloning, characterization, and expression of the put genes in the presence of root exudates. J Bacteriol 182: 91–99. Villacieros M et al. (2005) Polychlorinated biphenyl rhizoremediation by Pseudomonas fluorescens F113 derivatives, using a Sinorhizobium meliloti nod system to drive bph gene expresssion. Appl Environ Microbiol 71: 2687–2694. Walton BT, TA Anderson (1990) Microbial degradation of trichloroethylene in the rhizosphere: potential application to biological remediation of waste sites. Appl Environ Microbiol 56: 1012–1016. Wang K, K Conn, G Lazarovits (2006) Involvement of quinolinate phosphoribosyl transferase in promotion of potato growth by a Burkholderia strain. Appl Environ Microbiol 72: 760–768. White C, A Quiñones-Rivera, M Alexander (1998) Effect of wetting and drying on the bioavailability of organic compounds sequestered in soil. Environ Toxicol Chem 17: 2378–2382. Widdowson MA, S Shearer, RK Andersen, JT Novak (2005) Remediation of polycyclic aromatic hydrocarbon compounds in groundwater using poplar trees. Environ Sci Technol 39: 1598–1605. Wiesner MR, MC Grant, SR Hutchins (1996) Reduced permeability in groundwater remediation systems: role of mobilized colloids and injected chemicals. Environ Sci Technol 30: 3184–3191. Wilken A, C Bock, M Bokern, H Harms (1995) Metabolism of PCB congeners in plant cell cultures. Environ Toxicol Chem 14: 2017–2022. Wilson MS, EL Madsen (1996) Field extraction of a transient intermediary metabolite indicative of real time in situ naphthalene biodegradation Environ Sci Technol 30: 2099–2103. Wolfaardt GM, JR Lawrence, RD Robarts, DE Caldwell (1995) Bioaccumulation of the herbicide diclofop in extracellular polymers and its utilization by a biofilm community during starvation. Appl Environ Microbiol 61: 152–158. Yadav JS, RE Wallace, CA Reddy (1995) Mineralization of mono- and dichlorobenzenes and simultaneous degradation of chloro- and methyl-substituted benzenes by the white-rot fungus Phanerochaete chrysosporium. Appl Environ Microbiol 61: 677–690. Yao J, C Allen (2006) Chemotaxis is required for virulence and competitive fitness of the bacterial wilt pathogen Ralstonia solanacearum. J Bacteriol 188: 3697–3708. Ye D, MA Siddiq, A Maccubbin, S Kumar, HC Sikka (1996) Degradation of polynuclear aromatic hydrocarbons by Sphingomonas paucimobilis. Environ Sci Technol 30: 136–142. Yee DC, JA Maynard, TK Wood (1998) Rhizoremediation of trichloroethyelene by a recombinant, rootcolonizing Pseudomonas fluorescens strain expressing ortho-monooxygenase constitutively. Appl Environ Microbiol 64: 112–118.

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Evaluation of Bioremediation

INTRODUCTION An essential component of any bioremediation program is the evaluation of its success. Evaluation is not as readily accomplished as might be assumed for a number of reasons (Madsen 1991; Heitzer and Sayler 1993; Macdonald and Rittman 1993; Shannon and Unterman 1993). A valuable and critical review (Dott et al. 1995) discussed the need for an assessment that takes into account all the phases involved, and suggested that the difficulty of demonstrating in situ bioremediation may require long-term studies. An illustrative example is provided by bioremediation of an aquifer in Switzerland, which was contaminated with diesel fuel. Oxidants were provided in the form of O2 and NO3–, but measurements of carbon and nitrogen balance were not sufficiently precise to assess the spatial effectiveness of the treatment. A laboratory microcosm using material from the aquifer was constructed: the use of gene probes revealed that bacteria belonging to different phylogenetic groups were present, and bacteria able to degrade toluene and 1,3-dimethylbenzene were isolated (Hess et al. 1997). It was, however, pointed out that, because of the greater diversity of organisms in the aquifer, there were limitations in the extent to which the results of the laboratory experiments could be extrapolated to the field situation. Cardinal issues for assessing the effectiveness of bioremediation include the following: 1. Diminution in the concentration of the substrate(s) is not alone an acceptable measure of degradation, since loss may occur by volatilization or by transformation with the formation of transient or terminal metabolites. In addition there may be a continuous input, for example, by leaching. 2. The terrestrial environment is highly heterogeneous, so that representative sampling generally presents a serious problem, and the evaluation of spatial effectiveness is difficult (Hess et al. 1997). 3. The extent of leaching into groundwater may be difficult to evaluate, because few of the systems are sufficiently enclosed to make a convincing balance of the concentrations of the substrates and their metabolites—including CO2 or CH4 —that are lost to the atmosphere. 4. Most field environments are dynamic. There are therefore continuous changes in the input and in the metabolites that are being produced. Interpretation of chemical analyses may therefore be equivocal, and metabolites may be removed at different rates. One approach has been to use surrogate contaminants, for example, (E/Z)-chlorofluoroethene for chloroethene, in which only the E isomer is dechlorinated (Ennis et al. 2005). 5. Chemical analysis alone is insufficient in view of analytical uncertainties in the optimal methods for extraction. These must take into account both the free and associated analytes (Cajthaml and Sasek 2005), as well as metabolic products. A number of strategies have been proposed as potentially attractive solutions to evaluate the number of specific degradative organisms, or of the appropriate degradative enzymes. These procedures include chemical analysis of contaminants and degradation products, the analysis of populations, and the use of stable isotope enrichment, which are discussed below. All of them require knowledge of the degradation pathways under the relevant conditions, and this is generally based on laboratory studies with pure or mixed cultures achieved by enrichment at the expense of the contaminant. Although many of the examples that are given refer to laboratory experiments, the principles that

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emerge are relevant to field application with appropriate changes in the methodology. The questions all of them attempt to address are (a) is degradation taking place? and (b) at what rate?

ANALYSIS OF POPULATIONS Whereas in aquatic systems, bacteria are generally considered to be the major agents of biodegradation, the role of fungi in terrestrial systems may be equal or greater. Among fungi, considerable attention has been devoted to the metabolically versatile white-rot fungi (Shah et al. 1992; Barr and Aust 1994), and several examples are cited later. It is important to appreciate the wide range of organisms belonging to this group, even though greatest attention has hitherto been devoted to Phanerochaete chrysosporium. Further comments are given in Chapter 9, Parts 1 and 2, and the following discussion is restricted to bacteria. Pathways for the degradation of synthetic substrates have been elucidated in laboratory experiments using cultures of bacteria that have been obtained by elective enrichment with the given or a related substrate. This may, however, introduce a bias, not only in the organisms but also in the degradative genes that they contain (Marchesi and Weightman 2003). In addition, there are many organisms in natural habitats that have not hitherto been cultivated (Stevenson et al. 2004), and these will almost certainly play an important role in degradation. Natural enrichment certainly takes place in environments that have been chronically exposed to contaminants, and this is an important mechanism for the loss of agrochemicals. At the same time, bacteria that have been isolated from pristine environments with no established exposure to the contaminants may be able to degrade contaminants (e.g., Kitagawa et al. 2002; Sánchez et al. 2004). A number of procedures have been used for the analysis of natural populations including some, which are directed to specific metabolic activity, and use a variety of methods that have been developed in molecular biology. In general, these presuppose knowledge of the mechanisms for degradation of the contaminants, and this is equally true for methods using stable isotope analysis. In spite of the caution that has been directed to the limitations of pure culture methods, it is the use of these that has built up the background for these studies, which are not dependent on isolation of specific organisms. Application of polymerase chain reaction (PCR) to specific degradative organisms has been reviewed (Steffan and Atlas 1991).

ANALYSIS OF DEGRADATIVE POPULATIONS There has been substantial interest in identifying bacteria in environmental samples with a view to determining the components of bacterial communities and understanding their role in biodegradation and biotransformation. This has been possible by using molecular procedures including appropriate genetic probes to determine the existence of specific metabolic activity, and analysis of 16S rRNA and 16S rDNA to assess the similarity of isolates to other taxa. A valuable overview of the principles and application of a range of molecular procedures has been given (Power et al. 1998). Potential applications include: a. Analysis of natural degradation in the context of bioremediation that may usefully be combined with the use of stable isotope enrichment factors. b. Comparison of the degradation potential of natural populations with that of organisms isolated by enrichment. c. Prospecting for novel enzyme activity by analysis of soil metagenomes. These studies have revealed a number of cardinal aspects of microbial ecology: (a) the complexity of bacterial ecosystems, (b) the number of organisms of hitherto undetermined taxonomy, (c) the existence of organisms that have not been cultivated, and (d) the difficulty of assigning metabolic roles to many of the genomic sequences that have been revealed. Some examples are given to

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illustrate the range and application of these procedures. These include dioxygenation of aromatic compounds, alkane degradation, anaerobic dechlorination, and the discovery of hydrolytic enzymes for natural polymers.

APPLICATION TO THE DEGRADATION OF SPECIFIC CONTAMINANTS Hydrocarbons 1. Analysis of the occurrence and frequency of genes encoding the dioxygenases for aromatic hydrocarbons have been used to evaluate the potential of a site for bioremediation: a. An evaluation of the effect of plants on the remediation of a PAH-contaminated combined mineralization of 13C-labeled naphthalene in the bulk soil and in the rhizosphere, with the occurrence of genes specific for dioxygenation of naphthalene (ndoB), alkane monooxygenation (alkB), and catechol 2,3-dioxygenase (xylE) (Siciliano et al. 2003). The results showed that even aged contaminants were susceptible to plant-mediated degradation in the rhizosphere. b. A study directed to the analysis of aromatic hydrocarbon degradation used real-time PCR amplification of oxygenase genes. The primer sets were identified for the large subunits (α) of genes for dioxygenases including those for naphthalene, biphenyl, and toluene, and monooxygenases for xylene, phenol, and ring monooxygenases for toluene (Baldwin et al. 2003). c. The abundance of genes similar to the α-subunit of the nagAc naphthalene dioxygenase gene in Ralstonia sp. strain U2, which encodes the degradation of naphthalene to salicylate, was analyzed by real-time PCR in samples from a coal tar–contaminated site (Dionisi et al. 2004). There was a strong positive correlation between the number of gene copies and the concentration of naphthalene at the site, and a lower correlation with concentrations of phenanthrene, and it was concluded that active biodegradation was occurring. d. Isolates from a site contaminated with benzene, toluene, ethylbenzene, and xylenes (BTEX) were analyzed to assess the diversity of dioxygenases in the α-subunit of the toluene/biphenyl subfamily and the ability to grow with benzene, toluene, and ethylbenzene. All strains could degrade benzene, but only a few of them could degrade toluene and ethylbenzene. There was strong correlation between sequence type and substrate utilization, and sequence analysis showed that dioxygenases belonging to the isopropylbenzene is branch-dominated (Witzig et al. 2006). 2. A reverse sample genome probe procedure has been used to monitor sulfate-reducing bacteria in oil-field samples (Voordouw et al. 1991). It was extended to include 16 heterotrophic bacteria (Telang et al. 1997), and applied to evaluating the effect of nitrate on an oil-field bacterial community. Denatured genomic DNAs were used in a reverse sample genomic probe procedure to examine the effect of toluene and dicyclopentadiene on the community structure of bacteria in a contaminated soil. Hybridization of total community DNA isolated from soil exposed to toluene showed enrichment of strains that were able to metabolize toluene. At the same time, DNA from the soil contaminated with dicyclopentadiene indicated the enrichment of organisms that were able to form products by oxygenation with masses of 146 and 148, which corresponded to derivatives with epoxide or ketone groups. In contrast, no mineralization of dicyclopentadiene was observed (Shen et al. 1998). 3. The archaeal population in anaerobic oil-contaminated water at a crude oil storage site was examined using PCR amplification of fragments of 16S rDNA using a range of primers. Although two of the sequences were closely related to Methanosaeta concilii and Methanomethylovorans hollandica, other sequences could not be identified, including the most abundant (KuA12) that accounted for ≈50% of the total archaeal rDNA copies that

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were detected. Species of novel archaea were enriched at the site, and were putatively active methanogens (Watanabe et al. 2002). 4. A stable mixed culture that could degrade benzene was established, and the specific role of sulfate-reducing bacteria was inferred from the inhibition of degradation in the presence of molybdate. The consortium was characterized from the sequence of small subunit rRNA genes, some of which could be assigned to the family Desulfobacteriaceae (Phelps et al. 1998). Trichloroethene 1. Populations of bacteria had been enriched by adding lactate to enrich the indigenous organisms for in situ dechlorination of trichloroethene (TCE) to ethene. The consortia in the groundwater source were assessed using PCR amplification of 16S rRNA genes. In a clone library, a homoacetogen dominated the archaea, with a clone affiliated to the acetoclastic methanogen Methanosaeta concilii, while proteobacteria similar to Geobacter (Trichlorobacter) thiogenes and Sulfurospirillum multivorans were also found. A mixed culture that dechlorinated TCE to ethene was enriched from the groundwater and this was compared with that in the groundwater by terminal restriction fragment length polymorphism (RFLP). Although the enrichment community was less diverse than that from the groundwater, its archaeal structure was similar. It was concluded that acetate produced from the fermentation of lactate was the source of methane by the acetoclastic pathway (Macbeth et al. 2004). 2. The use of toluene to induce oxygenation of haloalkanes has been discussed in Chapter 7, Part 3, and probes for toluene-2-monooxygenase have been used to evaluate the potential number of TCE-degrading organisms in an aquifer (Fries et al. 1997b). In this study, repetitive extragenic palindromic PCR (REP-PCR) (de Bruijn 1992) of isolates was used to classify their metabolic capability. 3. To assess the population of organisms capable of degrading TCE at a contaminated site, samples were examined from a contiguous, but pristine site. Attention was directed to methane- and ammonia-oxidizing bacteria, and included both free-living and attached communities. Primer sets included genes for particulate methane monooxygenase and ammonia monooxygenase. Whole-genome amplification was used for metagenomic DNA from each of the communities at the sampling sites. Bacteria related to Methylocystis dominated both the free-living and attached communities, and gene sequences associated with TCE metabolism were observed for both of them (Erwin et al. 2005). Phenol 1. Analysis of the populations of two phenol-degrading bacteria Pseudomonas putida BH and Comamonas sp. strain E6 introduced into an activated sludge system was carried out by extraction of DNA, PCR amplification of the gyrB gene fragments using strain-specific primers, and quantification after electrophoresis by densitometry. Appropriate dilutions of the extracted DNA were used to maintain linearity between the measured intensity and population density. The initial inocula of ca. 108 cells/mL fell to ca. 104 cell/mL after 10 d and thereafter remained steady for a further 18 d (Watanabe et al. 1998a). 2. A study of the community in a phenol-degrading population used amplification of the gene encoding the largest subunit of phenol hydroxylase, followed by analysis using temperature gradient gel electrophoresis. Bacteria were also isolated by three procedures: (i) direct plating on a complex medium or a mineral medium supplemented with phenol, (ii) by enrichment in batch culture, and (iii) by chemostat enrichment. The dominant population could be isolated only by direct plating or by chemostat enrichment. One of the dominant organisms that contained a novel large phenol hydroxylase subunit was closely related to

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Variovorax paradoxus, and it was proposed that this was the principal organism for degradation of phenol in the community (Watanabe et al. 1998b). 3. A site at the Agricultural Experimental Station (Ithaca, NY) was treated in microcosms with 13C-labeled glucose, phenol, caffeine, and naphthalene. Levels of 13CO2 were measured to assess utilization of the substrates, and the populations analyzed by separating the 13 C-labeled DNA by density centrifugation, followed by PCR amplification and sequencing of 16S rRNA (Padmanabhan et al. 2003). Populations contained relatives to a range of bacteria that varied with the substrate. Only relatives of Acinetobacter were found in all samples, and for caffeine only Pantoea. 4. In a further study (DeRito et al. 2005) at the same site, attention was concentrated on phenol using respiration of 13C-labeled phenol and stable isotope probes of soil DNA. It was shown that the whole community could utilize products from the degradation of phenol. The distribution of labeled carbon was determined by using different regimes for the addition of substrate, and populations could be distinguished on the basis of their diversity. The greatest diversity was found in organisms from soil that had not been enriched with phenol, while members of the Gram-positive genera Kocuria and Staphylococcus dominated the phenol-degrading population. Members of the genus Pseudomonas dominated those that were metabolically versatile and were able to utilize the carbon from other organisms. Chlorophenol Soil samples were enriched in medium containing 2,4,6-trichlorophenol and subjected to a procedure to reduce the complexity of the mixed flora. This involved cyclic serial dilution, plating, and growth in liquid medium. REP-PCR and amplified ribosomal DNA restriction analysis (ARDRA) were used to monitor the flora during the cycles. From the fourth cycle, four organisms were isolated. Although three of them were unable to utilize 2,4,6-trichlorophenol, an additional slow-growing organism that was able to do so was isolated, and tentatively assigned to the genus Nocardioides (Maltseva and Oriel 1997). Chlorobenzoate The occurrence of bacteria that can mineralize 3-chlorobenzoate has been examined in soil samples from widely separated regions in five continents (Fulthorpe et al. 1998). The genotypes of the isolates were examined by two procedures: (i) REP-PCR genomic fingerprints (de Bruijn 1992) and (ii) analysis of restriction digests of the 16S rRNA (ARDRA) (Weisberg et al. 1991). The results showed that each genotype was generally (91%) restricted to the site from which the samples were collected, and that therefore the genotypes were not derived by global dispersion. Phenylurea Herbicides 1. DNA was extracted from soil samples that had been treated with three phenylurea herbicides during 10 years and from an untreated control. The PCR products using two primers were analyzed by denaturing gradient gel electrophoresis (DGGE), and the patterns used to assess the quantitative similarities of the bands. PCR using two different primers followed by DGGE was used to obtain 16S rDNA for sequencing (El Fantroussi et al. 1999). The microbial diversity determined from the gel profiles had decreased in the treated soils, and the sequencing revealed that the organisms that were most affected belonged to hitherto uncultivated taxa. Enrichment cultures showed that dichlorinated linuron was more readily degraded in cultures using the treated soils than from those using the untreated soils. DGGE analysis and sequencing showed that one of the components, which was found only in these enrichments, showed a 95% similarity to Variovorax sp. that was also found in enrichments using 2,4-dichlorophenoxyacetate (Kamagata et al. 1997). The

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effect of the herbicide applications had therefore affected both the composition of the bacteria flora and the metabolic capabilities of its components. 2. Analysis of a bacterial population that could degrade the chlorinated phenylurea herbicide linuron revealed the presence of several strains of bacteria, which were identified on the basis of 16S rRNA gene sequences, isolation of the strains, and determination of their metabolic role (Dejonghe et al. 2003). Five strains were identified of which Variovorax strain WDL1 was able to use linuron as a source of carbon, nitrogen, and energy. Among the other strains, Hyphomicrobium sulfonivorans was able to degrade N,O-dimethylhydroxylamine, while degradation of linuron was stimulated by the presence of these other strains. Dehalogenation of Chloroalkanoates Dehalogenases from bacteria enriched from activated sludge for the degradation of 2,2-dichloropropionate were compared with those from the environmental metagene pool. Although the dehalogenases found in pure cultures dominated the enrichment culture, they were only a minor part of the community used for enrichment. Analysis of dehalogenase genes found in pure cultures with those from the metagenome pool from which they were isolated revealed the substantial bias introduced by culturing both in the bacteria and in their degradative genes (Marchesi and Weightman 2003). Dechlorination of PCBs Metabolic inhibitors were used in an elegant set of experiments designed to enrich for the anaerobic organisms that carried out the ortho dechlorination of 2,3,5,6-tetrachlorobiphenyl (Holoman et al. 1998), and the community structure was followed by analysis of total community genes for 16S rRNA. It was shown that the diversity of the community could be reduced in mineral medium by the addition of inhibitors for methanogens (2-bromoethanesulfonic acid) and Clostridium spp. (vancomycin), without eliminating dechlorination that was inhibited by the addition of molybdate, which inhibited sulfate reduction. The bacteria that actively carried out ortho dechlorination belonged to three groups: the δ group, the low G + C Gram-positive group, and the Thermotogales subgroup that had not hitherto been implicated in anaerobic dechlorination.

APPLICATION TO SPECIFIC GROUPS OF ORGANISMS There has been considerable interest in the anaerobic metabolism of methane in the large reservoirs that lie beneath the seafloor, since little of this reaches the oxic conditions in the water column. Consortia of archaea that have so far resisted isolation and sulfate-reducing bacteria have been implicated (Orphan et al. 2002): 1. A hydrothermal vent in the Gulf of California produces hydrocarbons that support the growth of sulfate-reducing bacteria. PCR-amplified genes for sulfite reductase (dsrAB) and 16S rRNA were analyzed in sediment cores. Groups related to Desulfobacter that could utilize acetate represented a major group, while organisms affiliated with members of the genus Desulfotomaculum and organisms in the δ proteobacteria were also found (Dhillon et al. 2003). 2. Cores from the same vent were examined by 16S rRNA sequencing complemented with isotopic analysis of lipids. Uncultured groups associated with anaerobic methane oxidation ANME-1 and ANME-2 were identified, while δ13C values for archaeal lipids (archaeol, sn-2-hydroxyarcheol, and diphytanediol) indicated their origin in methanotrophic archaea (Teske et al. 2002). Analysis of the sequences for methyl coenzyme M reductase A (mcrA) in ANME-1 and ANME-2 suggested its potential catalytic activity (Hallam et al. 2003). It was proposed that anaerobic methane oxidation could occur by the reversal of methanogenesis, since many, though not all, of the genes could be identified in both pathways (Hallam et al. 2004).

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NONDIRECTED EXAMINATION OF NATURAL POPULATIONS 1. Core samples from three redox zones at depths of 6.1, 7.3, and 9.2 m in an aquifer contaminated with hydrocarbons and chlorinated solvents were examined (Dojka et al. 1998). Small subunit rRNA genes from core DNA were amplified directly by PCR with Bacteriaor Archaea-specific primers and cloned, and the clones screened by RFLP. Analysis of the sequences showed the existence of the following types: (a) 10 with no known taxonomic divisions, (b) 21 with no cultured representatives, and (c) 63 with recognized divisions. The sequences were classified into seven major groups, and it was shown that the two most abundant sequence types could be correlated with sequences of Methanosaeta sp. and Syntrophus sp. It was proposed that hydrocarbon degradation proceeded by oxidation of the hydrocarbons, fermentation of the resulting carboxylic acids by organisms with sequences related to Syntrophus sp., followed by acetoclastic methanogenesis by organisms related in sequence to Methanosaeta sp. 2. Small subunit rRNA genes were amplified directly by PCR in DNA from a sediment sample in Yellowstone National Park and cloned: universally conserved or Bacteriaspecific rDNA primers were used (Hugenholtz et al. 1998). RFLP was used to classify rDNA fragments. Most of the sequences were representatives of established bacterial division, although 30% were unrelated. Database matches suggested the presence of organisms closely related to (a) Thermodesulfovibrio yellowstonii that could carry out anaerobic sulfate reduction, (b) the organotrophs Thermus sp. and Dictyoglomus thermophilum, and (c) the hydrogen-oxidizing Calderobacterium hydrogenophilum. In contrast to previous perceptions, however, members of Bacteria dominated over Archaea in this sediment, so that the ecological boundaries between Bacteria and Archaea were less clearly defined. 3. The organisms in a high-salinity evaporation pond were examined by plating on solid media with prolonged incubation, and by analysis of a library of PCR-amplified 16S rRNA genes (Burns et al. 2004). The isolates were related to species of several genera including Haloferax, Halorubrum, and Natromonas. The first of these was not, however, represented in the gene library, and it was suggested that this could be the result of the ability to form colonies, even when it was not a dominant group. A major group identified in the library was the SHOW square organisms that have so far defied cultivation.

APPLICATION OF STABLE ISOTOPES The application of stable isotopes has been introduced in Chapter 6, and the following is devoted to illustrating specific applications.

STABLE ISOTOPE PROBES This procedure makes it possible to relate function in the environment to specific organisms or enzymes that carry out degradation. 1. Application to a site contaminated with coal tar waste and containing aromatic hydrocarbons was examined using a number of procedures. These included the use of 13C-labeled naphthalene followed by fractionation of labeled DNA, and construction of a library of cloned bacterial 16S rRNA. These were complemented with conventional assays for respiration, and isolation of bacteria by serial dilution and plating on a mineral medium during exposure to naphthalene. A strain CJ2 that closely resembled Polaromonas vacuolata was isolated from the contaminated sediment and was active in situ, even though the naphthalene dioxygenase gene was not related to that of existing strains of the genus (Jeon et al. 2003). 2. A phenol-degrading community was examined using 13C phenol followed by analysis of the stable-isotope-labeled RNA by equilibrium density centrifugation, and complemented

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with reverse-transcription-PCR and DGGE. The results suggested the dominance of a species of Thauera in addition to organisms conventionally associated with phenol degradation (Manefield et al. 2002). a. The population of putative methanotrophs in a peat soil microcosm was examined by long-term incubation with 13CH4 followed by separation of DNA by gradient centrifugation. This was used as a template for PCR. The amplified products of 16S rRNA genes and those for genes that encode enzymes established for methane oxidation were analyzed (Morris et al. 2002). Although sequences related to type I and type II methanotrophs were identified, the existence of clones related to the β-subclass of Proteobacteria indicated the presence of other groups. b. In an analogous study, the assimilation of 13CH3OH or 13CH4 was examined (Radajewski et al. 2002). Analysis showed the presence of bacteria belonging to the genera Methylocella, Methylocystis, and Methylocapsa, while DNA from the microcosms that was used as a template for PCR with ammonia oxygenase-specific primers showed sequences belonging to Nitrosomomas and Nitrosospira. A number of other sequences were observed, although their function was not resolved. c. The methylotrophic community in soil from rice paddies was examined by a number of methods including 13C labeling with methanol followed by incubation under aerobic conditions and analysis of 13C-labeled rRNA by density measurement in a Cs trifluoroacetate gradient. Sequence analysis of labeled rRNA revealed the presence of bacteria belonging to the Methylobacteriaceae, although with increasing length of incubation the Methylophilaceae dominated. In addition, labeled nucleic acids were observed in fungi and protozoans, and it was concluded that the 13C accumulated in the methylotrophs had been transferred during incubation to these organisms (Lueders et al. 2004a). Gasoline-contaminated groundwater containing BTEX from an aquifer was supplemented with [13C]benzene and an electron acceptor. Analysis of the 13C-RNA fraction revealed the presence of a phylotype related to the genus Azoarcus only when nitrate was used as supplement. This was confirmed by DGGE of 16S rRNA gene fragments. Isolation on a nonselective medium and screening of the colonies by DGG resulted in the isolation of two strains that were able to degrade benzene, toluene, and m-xylene though—not o- or p-xylene. It was suggested that the success of this procedure lay in combining stable isotope probing with DGGE screening of strains isolated on nonselective medium (Kasai et al. 2006). It has been established that methane is produced on rice roots by reduction of CO2. This was examined in rice roots using a combination of 16S rRNA sequencing and density gradient fractionation of 13C-labeled DNA after incubation with 13CO2. The major groups of archaea detected were Methanosarcinaceae that decreased with time to be replaced by the hitherto uncultured Rice Cluster I, although the former subsequently dominated (Lu et al. 2005). The syntrophic population of anaerobic organisms that carry out the oxidation of propionate was examined using [13C]propionate. Analysis of the “heavy” 13C-labeled rRNA by terminal restriction length polymorphism fingerprinting and sequence analysis of the resolved populations showed the presence of the genera Syntrophobacter, Smithella, and Pelotomaculum whose degradation of propionate had already been confirmed in other laboratory studies. Among the archaea, species of Methanobacterium, Methanosarcina, and members of the Rice Cluster I were found (Lueders et al. 2004b).

APPLICATION OF STABLE ISOTOPE ENRICHMENT This has attracted considerable attention in evaluating the effectiveness of remediation either by natural populations or those enriched by specific additions. Isotopic analysis of the contaminant supplemented with that of established metabolites makes it possible to determine whether degradation

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and transformation has taken place within the impacted area. It may also be possible to determine the rates at which this has taken place, and sometimes degradation pathways. Results from these experiments have been reported in a number of ways depending on the objective, and these have been summarized in Chapter 6. Illustrative examples include the following. Application of δ13C Chlorinated Aromatic Hydrocarbons 1. Values of δ13C have been measured for a number of polychlorinated biphenyl (PCB) congeners and applied to a number of commercial PCB mixtures. Both the number and the position of the chlorine substituents affected the depletion of 13C, and this reflected the manufacturing procedures that involved kinetic isotope effects as well as the source of the biphenyl starting material (Jarman et al. 1998). It was suggested that this could be applied to determine the source of PCBs in the environment. 2. An anaerobic bacterial enrichment culture was used to examine the dechlorination of 2,3,4,5-tetrachlorobiphenyl that produced 2,3,5-trichlorobiphenyl exclusively. Although there was no alteration in the values of δ13C, compound-specific analysis of Arochlor 1268 showed that there was a trend for decreasing 13C abundance with increasing content of chlorine. This is consistent with dechlorination of the congeners with more chlorine substituents. It was suggested that reductive dechlorination of PCBs would produce congeners with more depleted values of δ13C compared with unweathered Arochlors (Drenzer et al. 2001). 3. Values of δ13C were measured for the degradation of trichlorobenzenes (Griebler et al. 2004). For aerobic degradation by Pseudomonas sp. strain P51, which carried out degradation by dioxygenation, the values were not significant. In contrast, the isotope enrichment factor (ε) for anaerobic dechlorination by Dehalococcoides sp. strain CDB1, which produced 1,3-dichlorobenzene from 1,2,3-trichlorobenzene, was –3.4 ppm, and for 1,2,4trichlorobenzaene, which produced 1,4-dichlorobenzene, was –3.2 ppm. Methyl tert-Butyl Ether Enrichment factors have been used in studies with methyl tert-butyl ether (MTBE) that is a fuel additive and is a widespread contaminant in subsurface aquatic systems. 1. In laboratory enrichments with MTBE under both methanogenic and sulfate-reducing conditions, carbon enrichment of –14.4 ± 0.7 ppm was found (Somsamak et al. 2006), and in a methanogenic enrichment –14 ± 4.5 ppm for tert-amyl methyl ether (Somsamak et al. 2005). 2. Both δ13C and δ2H were used in an examination of several contaminated groundwater sites (Kuder et al. 2005). The isotopes were extensively fractionated during anaerobic degradation with stable enrichment factors of –13 ppm for carbon and –16 ppm for hydrogen. These values were used to predict the extent of biodegradation at the sites using the Raleigh model for degradation. 3. In a field investigation of groundwater where there were several sources of MTBE, values of δ13C ranged from –26 to +40 ppm, and δ2H from –73.1 to +60.3 ppm that were consistent with degradation along the plume (Zwank et al. 2005). At the same time, the carbon composition of tert-butanol that is produced by demethylation was constant, which showed that this was not further degraded. Comparison with previous measurements and important aspects of the methodology were discussed. Aromatic Hydrocarbons Investigations have examined applications to monocyclic aromatic hydrocarbons and phenols, under both aerobic and anaerobic conditions.

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Aerobic Conditions

1. Two strains for which the mechanism of biodegradation of benzene had not been established were examined for carbon and hydrogen isotope fractionation (Hunkeler et al. 2001). Values for Acinetobacter sp. and Burkholderia sp. were, respectively, –1.46 and –3.53 ppm for carbon and –12.8 and –11.2 ppm for hydrogen. Values of δ13C and δ2H for benzene differed from various sources, and it was pointed out that these differences should be taken into consideration. 2. Fractionation factors (ε) were measured for the aerobic degradation of aromatic hydrocarbons using several strains of bacteria that used different mechanisms for its initiation (Morasch et al. 2002). For strains of Pseudomonas putida strain F1 and naphthalene (strain BCIB 9816) that degrade toluene by dioxygenation, values for toluene were not significant. In contrast, for strains that initiated the degradation of alkylbenzenes by monooxygenation, values for Pseudomonas putida strain mt-2 were –3.3 ppm for toluene, –1.7 ppm for m-xylene, and –2.3 ppm for p-xylene, and for Ralstonia pickettii strain PJO1 –1.1 ppm for toluene. Anaerobic Conditions

1. Under laboratory methanogenic and sulfidogenic conditions, only low values of enrichment factors were observed, although 13C enrichment of toluene was found at late stages of transformation (Ahad et al. 2000). 2. In a continuous flow headspace system, δ2H analysis was used during methanogenic degradation of toluene (Ward et al. 2000). Enrichment values exceeding –60 ppm were found, and these are about 10 times greater than the values reported for δ13C. It was suggested that combined use of both δ13C and δ2H would be attractive for assessing bioremediation at contaminated sites. 3. Four different conditions for the degradation of toluene were examined in cultures that used different electron acceptors (oxygen, nitrate, ferric iron, or sulfate), and different pathways for degradation (Meckenstock et al. 1999). The fractionation factors for Pseudomonas putida (1.0026), Thauera aromatica (1.0017), Geobacter metallireducens (1.0018), and a sulfate-reducing organism (1.0017) were in the same range for all the organisms. 4. Enrichment factors for both carbon and hydrogen were measured during anaerobic degradation of benzene using different electron acceptors (Mancini et al. 2003). There were no significant differences in the enrichment factors among the cultures, with values for δ13C from –1.9 to –3.6 ppm, and much larger values for δ2H (–29 to –79 ppm). This would be consistent with different mechanisms for degradation involving C–H bond fission. 5. Enrichment factors during the anaerobic degradation of o-xylene, m-xylene, m-cresol, and p-cresol by pure cultures of sulfate-reducing bacteria that use the fumarate pathway ranged from –1.5 to –3.9 ppm (Morasch et al. 2004). It was therefore proposed that this could be applied to evaluating in situ bioremediation of contaminants that use this pathway for biodegradation. 6. A sulfate enrichment culture prepared from a contaminated site gave enrichment factors (ε) of –1.1 for naphthalene and –0.9 for 2-methylnaphthalene (Griebler et al. 2004). These values combined with literature values from analogous laboratory experiments were used to quantify degradation of toluene, xylenes, and naphthalene at the site. Additional evidence for degradation of BTEX was derived from analyses of established metabolites produced by anaerobic degradation. 7. Direct evidence for the biodegradation of benzene and toluene in a contaminated aquifer was lacking, and an alternative strategy was examined. Bio-Sep beads were maintained in tubes and [13C] benzene or [13C]toluene were sorbed on to the surface. Analysis of δ13C in fatty acids extracted from lipids showed enrichments up to 13,500 ppm for benzene and

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toluene, and significant differences between values of individual fatty acids for the benzene and toluene amendments. It was therefore concluded that both benzene and toluene in the aquifer were being degraded, and in addition that different groups of organisms were responsible (Geyer et al. 2005). Chloroalkanes and Chloroalkenes Aerobic Conditions

1. For C1 and C2 chloroalkanes and chloroalkenes, δ13C values ranged from –25.58 (trichloroethane) to –58.77 (CH3Cl), and δ37Cl values from –2.86 (trichloroethane) to +1.56 (CH2Cl2). Although for analysis of environmental samples the method has the disadvantage that water must first be removed from the samples, it has been used to determine the distribution of TCE in a contaminated aquifer (Sturchio et al. 1998). 2. During the aerobic degradation of chloroethene (vinyl chloride [VC]) by strains of Mycobacteria and Nocardioides, enrichment factors (ε) lay within the range –8.2 ± 0.1 to –7.0 ± 0.3 ppm (Chartrand et al. 2005). These values were lower than those for the anaerobic degradation of chloroethene, and values of the kinetic isotope effect were 1.10 ± 0.001 for aerobic degradation, compared with the value of 1.03 ± 0.007 for anaerobic degradation. These are consistent with the pathways in which the aerobic reaction involves the C=C bond compared with the higher mass of the C–Cl bond in the anaerobic reaction. 3. Experiments have been carried out with 1,2-dichloroethane for which the pathways of degradation have been established. For degradation by Pseudomonas sp. strain DCA1 in which degradation is initiated by monooxygenation, the enrichment factor was –3.0 ppm, whereas for organisms including Xanthobacter autotrophicus in which a dehalogenase initiates degradation values of –32.3 ppm were found (Hirschorn et al. 2004). This procedure therefore differentiated the mechanisms by which degradation took place, and would be applicable to field studies. 4. Carbon isotope fractionation was examined during the aerobic degradation of TCE by Burkholderia cepacia strain G4 that possesses toluene monooxygenase activity (Barth et al. 2002). There were substantial differences in values of isotope shifts during degradation, from –57 to –17 ppm, and when the data were corrected to correspond to the same amount of substrate reduction the Releigh enrichment factor was –18.2. Anaerobic Conditions

1. A methanogenic enrichment culture was used to determine δ13C enrichment factors during the dechlorination of TCE, cis-1,2-dichloroethene (cis-DCE), and chloroethene (VC) (Bloom et al. 2000). Using the Raleigh model, enrichment factors (ε) were –6.6 and –2.5 ppm for TCE, –14.1 and –16.1 ppm for cis-DCE, and –26.6 and –21.5 ppm for VC. The large isotope effects that were observed suggested the value of this procedure for evaluating anaerobic dechlorination of chlorinated ethenes. 2. Anaerobic microbial consortia from different sources were used to determine enrichment factors for TCE, cis-dichloroethene, and chloroethene (VC) during dechlorination (Slater et al. 2001). For one of them (KB-1) that was able to bring about reductive dechlorination using methanol as electron donor, Raleigh enrichment factors values were –13.8, –20.4, and –22.4 ppm, whereas for tetrachloroethene for which the Raleigh model could not be applied, values varied among the consortia. It was suggested that this procedure could provide a valuable indication of dechlorination in spite of the rather wide spread of the enrichment factors that were observed. 3. A site at which earlier attempts at remediation of TCE did not proceed beyond production of cis-dichloroethene was augmented with the KB-1 culture, which had already been

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examined in microcosms. The isotope levels of cis-dichloroethene and chloroethene were increased at the final sampling, although that of the continuous input of TCE remained essentially constant (Chartrand et al. 2005). 4. A comparison had been made of 13C fractionation during the dechlorination of tetrachloroethene by Sulfurospirillum multivorans and Desulfitobacterium sp. strain PCE-S in laboratory experiments (Nijenhuis et al. 2005). Isotope fractionation in growing cultures was 1.0052 for Desulfitobacterium sp. and only 1.00042 for Sulfurospirillum multivorans, whereas fractionation was greater in crude cell extracts from both strains. It was concluded that caution should therefore be exercised in applying fractionation factors to the evaluation of in situ bioremediation. Isotopes of oxygen, sulfur, and chlorine have been much less extensively used. Illustrations include the following. Application of δ18O

The values for δ13C and δ18O were examined during degradation of diesel oil by a mixed culture under aerobic conditions. Oxygen in the gas phase of closed samples was analyzed by conversion into CO2 after cryogenic separation. The values for oxygen were particularly valuable in correlating the production of carbon dioxide with the loss in substrate concentration (Aggarwal et al. 1997), and it was therefore suggested that this methodology could be used to provide rates of in situ biodegradation. Application of δ34S

1. Determinations of δ34S have been extensively used in studies on the sulfur cycle, including reactions involving microbial anaerobic reduction of sulfate and thiosulfate (Smock et al. 1998). 2. Isotope fractionation during sulfate reduction by the hyperthermophilic Archaeoglobus fulgidus varied with the concentration of sulfate, and it was suggested that different pathways were operative at concentrations >0.6 or <0.3 mM (Habicht et al. 2005). 3. Enrichment factors with a range of anaerobic sulfate-reducing bacteria under optimal conditions yielded fractionation factors ranging from –2 to –42 ppm based on the isotopic composition of sulfate (Detmers et al. 2001). Although values for organisms that completely oxidized the substrate (benzoate, acetate, or butyrate) were generally greater than for those that carried out only the incomplete oxidation of lactate, the physiology and metabolic competence of the organisms clearly determined the extent of fractionation. Application of δ37Cl

1. Isotope ratios for 13C and 37Cl have been measured, and applied to chloroalkanes and chloroalkenes (Holt et al. 1997). For the C1 and C2 compounds, δ13C values ranged from –25.58 to –58.77, and δ37Cl values from –2.86 to +1.56. It was suggested that the method could be used to study the fate and distribution of such compounds. The method has, however, the disadvantage that in this method, water must first be removed from the sample. 2. Isotope ratios for 13C and 37Cl were measured for the aerobic degradation of dichloromethane by a methanotroph MC8b (Heraty et al. 1999). Values of the fractionation factor (α) were 0.9586 for carbon and 0.9962 for chlorine, and kinetic isotope effects were 1.0424 for carbon and 1.0038 for chlorine. 3. Three anaerobic dechlorinating consortia were used to examine fractionation during dechlorination of tetrachloroethene and TCE to cis-dichloroethene (Numata et al. 2002). Fractionation factors (α) for the first reaction ranged from 0.987 to 0.991 for the three consortia, and for the second reaction were 0.9944 for all consortia. Some important limitations were pointed out (a) the chlorinated ethenes were not separated so that the isotopic

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composition represented the totality of tetrachloroethene, TCE, and cis-dichloroethene and (b) the chlorine isotopic composition of the TCE that was used differed from the expected sum value for the chlorine atoms. 4. Kinetic isotope effects for the dechlorination by Xanthobacter autotrophicus strain GJ10 were 1.0045 for 1,2-dichloroethane, and 1.0066 for 1-chlorobutane (Lewandowicz et al. 2001). Collectively, these results illustrate the value of this procedure for determining the occurrence of biodegradation and biotransformation in natural environments, and their application to assessing the effectiveness of bioremediation. A number of important limitations should be addressed. a. Values may be strictly dependent on the pathway used for degradation, and this must be taken into consideration. b. For aromatic hydrocarbons and chloroethenes, large differences exist between reactions carried out under aerobic or anaerobic conditions, and between organisms carrying out anaerobic dechlorination. These should be carefully taken into account. c. For monocyclic aromatic hydrocarbons under aerobic conditions, values may depend on the mechanism of degradation, and the use of values for both 13C and 2H can provide valuable information. Combination with analyses of metabolites established from laboratory investigations provides a valuable complement. d. When mixtures of chlorinated alkenes are examined, it is essential to consider the range of possible mechanisms that could produce the products, since these may not necessarily reflect values for conventional sources (Hunkeler et al. 2005).

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Holoman TRP, MA Elberson, LA Cutter, HD May, KR Sowers (1998) Characterization of a defined 2,3,5,6tetrachlorobiphenyl-ortho-dechlorinating microbial community by comparative sequence analysis of genes coding for 16S rRNA. Appl Environ Microbiol 64: 3359–3367. Holt BD, NC Sturchio, TA Abrajano, LJ Heraty (1997) Conversion of chlorinated volatile organic compounds to carbon dioxide and methyl chloride for isotopic analysis of carbon and chlorine. Anal Chem 69: 2727–2733. Hugenholtz P, C Pitulle, KL Hershberger, NR Pace (1998) Novel division level bacterial diversity in a Yellowstone hot spring. J Bacteriol 180: 366–376. Hunkeler D, N Andersen, R Aravena, SM Bernasconi, BJ Butler (2001) Hydrogen and carbon isotope fractionation during aerobic biodegradation of benzene. Environ Sci Technol 35: 3462–3467. Hunkeler D, R Aravena, K Berry-Spark, E Cox (2005) Assessment of degradation pathways in an aquifer with mixed chlorinated hydrocarbon contamination using stable isotope analysis. Environ Sci Technol 39: 5975–5981. Jarman WM, A Hilkert, CE Bacon, JW Collister, K Ballschmiter, RW Risebrough (1998) Compound-specific carbon isotopic analysis of Arochlors, Clophens, Kanechlors, and Phenoclors. Environ Sci Technol 32: 833–836. Jeon CO, W Park, P Padmanabhan, C Derito, JR Snape, EL Madsen (2003) Discovery of a bacterium, with distinctive dioxygenase, that is responsible for in situ biodegradation in contaminated sediment. Proc Natl Acad Sci USA 100: 13591–13596. Kamagata Y, RR Fulthorpe, K Tamura, H Takami, LJ Forney, JM Tiedje (1997) Pristine environments harbor a new group of oligotrophic 2,4-dichlorophenoxyacetic acid-degrading bacteria. Appl Environ Microbiol 63: 2266–2272. Kasai Y, Y Takahata, M Manefield, K Watanabe (2006) RNA-based stable isotope probing and isolation of anaerobic benzene-degrading bacteria from gasoline-contaminated groundwater. Appl Environ Microbiol 72: 3586–3592. Kitagawa W, S Takami, K Miyauchi, E Masai, Y Kamagata, JM Tiedje, M Fukuda (2002) Novel 2,4-dichlorophenoxyacetic acid degradation genes from oligotrophic Bradyrhizobium sp. strain HW 13 isolated from a pristine environment. J Bacteriol 184: 509–518. Kuder T, JT Wilson, P Kaiser, R Kolhatkar, P Philp, J Allen (2005) Enrichment of stable carbon and hydrogen isotopes during anaerobic biodegradation of MTBE: microcosm and field evidence. Environ Sci Technol 39: 213–220. Lewandowicz A, J Rudzinski, L Tronstad, M Widersten, P Ryberg, O Matsson, P Paneth (2001) Chlorine isotope effects on the haloalkane dehalogenase reaction. J Am Chem Soc 123: 4550–4555. Lu Y, T Lueders, MW Friedrich, R Conrad (2005) Detecting active methanogenic populations on rice roots using stable isotope probing. Environ Microbiol 7: 326–336. Lueders T, B Pommerenke, MW Friedrich (2004b) Stable-isotope probing of microorganisms thriving at thermodynamic limits: syntrophic propionate oxidation in flooded soil. Appl Environ Microbiol 70: 5778–5786. Lueders T, B Wagner, P Claus, MW Friedrich (2004a) Stable isotope probing of rRNA and DNA reveals a dynamic methylotroph community and trophic interactions with fungi and protozoa in oxic rice field soil. Environ Microbiol 6: 60–72. Macbeth TW, DE Cummings, S Spring, LM Petzke, KS Sorenson (2004) Molecular characterization of a dechlorinating community resulting from in situ biostimulation in a trichloroethene-contaminated fractured basalt aquifer and comparison to a derivative laboratory culture. Appl Environ Microbiol 70: 7329–7341. Macdonald JA, BE Rittman (1993) Performance standards for in situ bioremediation. Environ Sci Technol 27: 1974–1979. Madsen EL (1991) Determining in situ biodegradation. Facts and challenges. Environ Sci Technol 25: 1663–1673. Maltseva O, P Oriel (1997) Monitoring of an alkaline 2,4,6-trichlorophenol-degrading enrichment culture by DNA fingerprinting methods and isolation of the responsible organism, haloalkaliphilic Nocardioides sp. strain M6. Appl Environ Microbiol 63: 4145–4149. Mancini SA, AC Ulrich, G Lacrampe-Couloume, B Sleep, EA Edwards, BS Lollar (2003) Carbon and hydrogen isotope fractionation during anaerobic biodegradation of benzene. Appl Environ Microbiol 69: 191–198. Manefield M, AS Whiteley, RI Griffiths, MJ Bailey (2002) RNA stable isotope probing, a novel means of linking microbial community function to phylogeny. Appl Environ Microbiol 68: 5367–5373.

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Marchesi JR, AJ Weightman (2003) Comparing the dehalogenase gene pool in cultivated α-halocarboxylic acid-degrading bacteria with the environmental matagene pool. Appl Environ Microbiol 69: 4375–4382. Meckenstock RU, B Morasch, R Warthmann, B Schink, E Annweiler, W Michaelis, HH Richnow (1999) 13 12 C/ C isotope fractionation of aromatic hydrocarbons during microbial degradation. Environ Microbiol 1: 409–425. Morasch B, HH Richnow, A Vieth, B Schink, RU Meckenstock (2004) Stable isotope fractionation caused by glycyl radical enzymes during bacterial degradation of aromatic compounds. Appl Environ Microbiol 70: 2935–2940. Morasch B, HH Richnow, B Schink, A Vieth, RU Mweckenstock (2002) Carbon and hydrogen stable isotope fractionation during aerobic bacterial degradation of aromatic hydrocarbons. Appl Environ Microbiol 68: 5191–5194. Morris SA, S Radajewski, TW Willison, JC Murrell (2002) Identification of the functionally active methanotroph population in a peat soil microcosm by stable-isotope probing. Appl Environ Microbiol 68: 1446–1453. Nijenhuis I, J Andert, K Beck, M Kästner, G Diekert, H-H Richow (2005) Stable isotope fractionation of tetrachloroethene during reductive dechlorination by Sulfospirillum multivorans and Desulfi tobacterium sp. strain PCE-S and abiotic reactions with cyanocobalamin. Appl Environ Microbiol 71: 3413–3419. Numata M, N Nakamura, H Koshikawa, Y Terashima (2002) Chlorine isotope fractionation during reductive dechlorination of chlorinated ethenes by anaerobic bacteria. Environ Sci Technol 36: 4389–4394. Orphan VL, CH Hpuse, K-U Hinrichs, KD McKeegan, EF DeLong (2002) Multiple archaeal groups mediate methane oxidation in anoxic cold seep sediments. Proc Natl Acad USA 99: 7663–7668. Padmanabhan P, S Padmanabhan, C DeRito, A Gray, D Gannon, JR Snape, DC Tsai, W Park, C Jeon, EL Madsen (2003) Respiration of 13C-labeled substrates added to soil in the field and subsequent 16S rRNA gene analysis of 13C-labeled soil DNA. Appl Environ Microbiol 69: 1614–1622. Phelps CD, LJ Kerkhof, LY Young (1998) Molecular characterization of a sulfate-reducing consortium, which mineralizes benzene. FEMS Microbiol Ecol 27: 269–279. Power M, JR van der Meer, R Tchelet, T Egli, R Eggen (1998) Molecular-based methods can contribute to assessment of toxicological risks and bioremediation strategies. J Microbiol Methods 32: 107–119. Radajewski S, G Webster, DS Reay, SA Morris, P Ineson, DB Nedwell, JI Prosser, JC Murrell (2002) Identification of active methylotroph populations in an acidic forest soil by stable-isotope probing. Microbiology (UK) 148: 2331–2341. Sánchez M Vásquez, B González (2004) A previously unexposed forest soil microbial community degrades high levels of the pollutant 2,4,6-trichlorophenol. Appl Environ Microbiol 70: 7567–7570. Shah MM, DP Barr, N Chung, SD Aust (1992) Use of white rot fungi in the degradation of environmental chemicals. Toxicol Lett 64: 493–501. Shannon MJR, R Unterman (1993) Evaluating bioremediation, distinguishing fact from fiction. Annu Rev Microbiol 47: 715–738. Shen Y, LG Stehmeier, G Voordouw (1998) Identification of hydrocarbon-degrading bacteria in soil by reverse sample genome probing. Appl Environ Microbiol 63: 637–645. Siciliano SD, JJ Germida, K Banks, CW Greer (2003) Changes in microbial community composition and function during a polyaromatic hydrocarbon phytoremediation field trial. Appl Environ Microbiol 69: 483–489. Slater GF, BS Lollar, BE Sleep, EA Edwards (2001) Variability in carbon isotope fractionation during biodegradation of chlorinated ethenes: implications for field applications. Environ Sci Technol 35: 901–907. Smock AM, ME Böttcher, H Cypionka (1998) Fractionation of sulfur isotopes during thiosulfate reduction by Desulfovibrio desulfuricans. Arch Microbiol 169: 460–463. Somsamak P, HH Richnow, MM Häggblom (2005) Carbon isotope fractionation during anaerobic biotransformation of methyl tert-butyl ether and tert-amyl methyl ether. Environ Sci Technol 39: 103–109. Somsamak P, HH Richnow, MM Häggblom (2006) Carbon isotope fractionation during anaerobic degradation of methyl tert-butyl ether under sulfate-reducing and methanogenic conditions. Appl Environ Microbiol 72: 1157–1163. Steffan RJ, RM Atlas (1991) Polymerase chain reactions, applications in environmental microbiology. Annu Rev Microbiol 45: 137–161. Stevenson BS, SA Eichorst, JT Wertz, TM Schmidt, JA Breznak (2004) New strategies for cultivation and detection of previously uncultivated microbes. Appl Environ Microbiol 70: 4748–4755.

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Sturchio NC, JL Clausen, IJ Heraty, L Huang, BD Holt, TA Abrajano (1998) Chlorine isotope investigation of natural attenuation of trichloroethene in an aerobic aquifer. Environ Sci Technol 32: 3037–3042. Telang AJ, S Ebert, LM Focht, DW Westlake, GE Jenneman, D Gevertz, G Voordouw (1997) Effect of nitrate injection on the microbial community in an oil field monitored by reverse sample genome probing. Appl Environ Microbiol 63: 1785–1793. Teske A, K-U Hinrichs, V Edgcomb, A de Vera Gomez, D Lysela, SP Sulva, ML Sogin, HW Jannach (2002) Microbial diversity of hydrothermal sediments in the Guaymas Basin: evidence for anaerobic methanotrophic communities. Appl Environ Microbiol 68: 1994–2007. Voordouw G, JK Voordouw, RR Karkhoff-Schweiser, PM Fedorak, DWS Westlake (1991) Reverse sample genome probing, a new technique for identification of bacteria in environmental samples by DNA hybridization, and its application to the identification of sulfate-reducing bacteria in oil field samples. Appl Environ Microbiol 57: 3070–3078. Ward JAM, JME Ahad, G Lacrampe-Couloume, GF Slater, EA Edwards, BS Lollar (2000) Hydrogen isotope fractionation of toluene: potential for direct verification of bioremediation. Environ Sci Technol 34: 4577–4581. Watanabe K, M Teramoto, H Futamata, S Harayama (1998b) Molecular detection, isolation, and physiological characterization of functionally dominant phenol-degrading bacteria in activated sludge. Appl Environ Microbiol 64: 4396–4402. Watanabe K, S Yamamoto, S Hino, S Harayama (1998a) Population dynamics of phenol-degrading bacteria in activated sludge determined by gyrB-targeted quantitative PCR. Appl Environ Microbiol 64: 1203–1209. Watanabe K, Y Kodama, N Hamamura, N Kaku (2002) Diversity, abundance, and activity of archaeal populations in oil-contaminated groundwater accumulated at the bottom of an underground crude oil storage cavity. Appl Environ Microbiol 68: 3899–3907. Weisberg WG, SM Barns, DA Pelletier, DJ Lane (1991) 16S ribosomal DNA amplification for phylogenetic study. J Bacteriol 173: 697–703. Witzig R, H Junca, H-J Hecht, DH Pieper (2006) Assessment of toluene/biphenyl dioxygenase gene diversity in benzene-polluted soils: links between benzene degradation and genes similar to those encoding isopropylbenzene dioxygenases. Appl Environ Microbiol 72: 3504–3514. Zwank L, M Berg, M Elsner, TC Schmidt, RP Schwarzenbach, SB Haderlien (2005) New evaluation scheme for two-dimensional isotope analysis to decipher biodegradation processes: application to groundwater contamination by MTBE. Environ Sci Technol 39: 1018–1029.

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INTRODUCTION This chapter is arranged according to the nature of the contaminants, and attempts to take into account the wide spectrum of compounds that have been deposited onto, or leaked into terrestrial sites, generally as a result of discontinued industrial activity. The discussion is enlarged to a discussion of groundwater that has been contaminated either directly or by leaching. Relevant experiments on bioremediation using a selection of procedures are used as illustration. In addition, experiments are discussed where results relevant to bioremediation have been obtained—even when application was not implemented. It should be emphasized that many aspects of the bioremediation of terrestrial sites impinge on the aquatic environment due to (a) leaching of compounds from the terrestrial environment, and (b) the facility of some contaminants with appreciable water solubility to be disseminated in groundwater. The general principles have been reviewed in Chapter 12, and some procedures for evaluating the effectiveness of remediation in Chapter 13. These contain illustrations of the application to many of the contaminants that are discussed in this chapter. Brief comments are given on the pathways of biodegradation to illuminate the discussion: further details may be found in Chapters 7 through 10.

PART 1 PETROLEUM HYDROCARBONS: REFINERY WASTE AND STRANDED OIL There are several different situations in which petroleum hydrocarbons pose a threat that has attracted solution by bioremediation. These include: (1) oil-refinery waste and contamination of the surrounding soil, (2) leakage from oil pipelines and underground storage tanks or basins, and (3) spillage of crude oil in the marine environment after accidents at sea. In important aspects, the basic issues of all of these are similar, and the strategy for experiments on bioremediation may be viewed against the background provided by a valuable review by Atlas (1981). The broad classes of compounds involved (alkanes, cycloalkanes, aromatic hydrocarbons, and heteroarenes) that are formed by transformation of plant and algal residues were summarized together with the range of microorganisms capable of carrying out their degradation. Important environmental determinants of biodegradation were outlined including the addition of nutrients, the concentration of oxygen, the temperature, the salinity, and the role of the hydrocarbon–water interface as the site of active degradation. It is worth pointing out that (a) the biodegradability of cycloalkanes has attracted less attention that the other components although some of them present potential recalcitrance, and (b) mass spectrometric methods using acetone chemical ionization (CI) have been developed to distinguish alkenes from their isomeric cycloalkanes (Roussis and Fedora 1997). Reviews on the microbial metabolism of hydrocarbons with biochemical aspects are available, and include those of Britton (1984) on alkanes, and of Morgan and Watkinson (1994) that also includes cycloalkanes and some aromatic compounds. Virtually all the issues that are discussed in these recur in the examples that are used as illustration. Some broad generalizations are summarized: 1. Aerobic degradation of hydrocarbons requires access to electron acceptors, generally oxygen in natural situations, added hydrogen peroxide in terrestrial systems, or nitrate or sulfate under anaerobic conditions that prevail at deeper levels of the soil or sediment. 2. The biodegradation of n-alkanes and the pathways whereby it is accomplished are well established. The range of compounds includes branched-chain compounds such as pristane 639

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and, although branching may present an obstacle (Pirnik 1977), this may be overcome by carboxylation (Rontani et al. 1997). Crude petroleum may also contain acidic derivatives of cycloalkanes (naphthenic acids) some of which are recalcitrant. The bioremediation of aromatic hydrocarbons is discussed in Part 2 of this chapter, and it is necessary only to recall that whereas naphthalene and phenanthrene are readily degradable, polycyclic aromatic hydrocarbons (PAHs) with more than five rings may be more recalcitrant. 3. Many bacteria produce surfactants in response to exposure to hydrocarbons, and these have been demonstrated both for those that degrade alkanes and PAHs (Déziel et al. 1996). The positive effect of adding surfactants is, however, equivocal (Deschênes et al. 1996). 4. Occasionally, biotransformation in the absence of substantial biodegradation may be acceptable. Dicyclopentadiene is produced pyrolytically in petrochemical plants and has a nauseating and penetrating odor. Although it could be degraded by a mixed bacterial culture, the major part was transformed into a series of oxygenated compounds that were presumed to be less malodorous (Stehmeier et al. 1996; Shen et al. 1998).

TERRESTRIAL HABITATS Alkanes A number of reviews cover different aspects of the bioremediation of soil contaminated with hydrocarbon refinery waste. The relative merits of in situ and on-site procedures, the role of electron acceptors and the addition of inorganic nutrients have been discussed by Dott et al. (1995). It was shown that inoculation even with organisms isolated from the sites—and putatively able to degrade the contaminants—had only very limited effect. The important determinants were therefore, provision of oxygen as O2 or H2O2, and of an inorganic source of nitrogen. This conclusion is supported by the results of attempts to improve the effectiveness of remediation of diesel-contaminated alpine soils by adding inorganic nutrients, or by inoculation with a psychrophilic culture of oil-degrading microorganisms (Margesin and Schinner 1997). The former was more effective, although a substantial loss occurred by abiotic processes, and only moderate remediation was achieved. The improvements induced by adding inorganic nutrients are, however, determined by a number of factors including the level of fertilization (Braddock et al. 1997). For example, the effect of spraying nitrate solutions to plots that had been contaminated with both BTEXTMB (TMB = trimethylbenzene) and jet fuel and from which the surface vegetation had been removed (Hutchins et al. 1998) was equivocal. Although hydrocarbon removal could be demonstrated, it was not possible to assign success solely to the addition of nitrate due to the intervention of soil washing. In microcosms prepared from core samples, however, removal occurred under nitrate and iron-reducing conditions, though not sulfate-reducing conditions. The modes of dissipation and a summary of remedial techniques including designs for solid-phase systems and bioreactors “biopiles” have been given (Bossert and Compeau 1995). It was concluded that the potential of bioremediation had not been fully exploited, and that the important issues of bioavailability and the procedures for evaluating effectiveness should be addressed. An investigation into the design of laboratory reactors for the degradation of diesel-contaminated soil illustrated several important factors: a. The advantage of periodic slurry systems with recycling at a rate that is determined by the degree of contamination of the soil. b. Periodic aeration was as effective as continuous aeration, and resulted in significantly less loss by volatilization (Cassidy and Irvine 1997). There is considerable interest in the removal of contaminants from the sites of former gasfilling stations. The surrounding soil is contaminated not only with fuel hydrocarbon residues, but also with alkyl lead compounds that were previously used as gasoline additives. It has been shown

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that while the latter are apparently degradable, this is inhibited by the presence of the concomitant hydrocarbons (Mulroy and Ou 1998). The presence of the additive MTBE in groundwater is discussed in a later section. Naphthenates These are found in crude petroleum including bitumen in the Athabasca tar sands of Northern Alberta. They contain a complex mixture of saturated polycyclic five- and six-membered cycloalkanes with alkane and alkanoic acid substituents. Attention has been directed to the degradation of both commercially available products, and those that are produced during bitumen extraction. Although the former were degradable (Clemente et al. 2004), the higher molecular mass components of the latter were much more recalcitrant (Scott et al. 2005).

MARINE HABITATS The Baffin Island Oil Spill Project The general aspects of the degradation of oil in the marine environment have been given by van der Linden (1978), who discussed important issues including the effect of adding nutrients and the significance of temperature. Attention is drawn to a unique investigation that involved the deliberate discharge of oil, and the use oil dispersants into a geographically remote and restricted environment in the Canadian Arctic. The reasons for adopting this strategy are important. It was felt that, in spite of the effort already devoted to the persistence and toxicity of hydrocarbons in the marine environment, much of this knowledge could not be validly extrapolated to an arctic environment with low water temperatures, and a restricted—and possibly—sensitive pristine biota. It is not realistic to review in detail the conclusions from this study, which have been published in a series of comprehensive articles in a supplement to Arctic (1987: 40) (Sergy 1987; Owens et al. 1987). This idea alone should be warmly welcomed as an alternative to the publication of details in miscellaneous reports that may not be readily accessible to interested members of the scientific community. This study is also important since, although it was designed to provide a predictive basis for future activity in the event of an oil spill, it also clearly illustrated the kinds of data that are required to provide the basis for epidemiological studies. Those aspects of its design that could profitably be incorporated into a wider range of impact assessment procedures are of most direct relevance in the present context, and an attempt is therefore made to summarize those features. 1. Comprehensive chemical analyses of samples of water, sediment, and biota were carried out both before and after the spill. This cannot of course be carried out in most cases, and illustrates a serious limitation in field studies, in which lack of background data or difficulty in finding an uncontaminated control locality is frequently encountered. Sum parameters were sparingly employed in Baffin Island Oil Spill (BIOS), and emphasis was placed on the analysis of specific compounds: attention was directed not only to PAHs, but also to azaarenes, dibenzothiophenes, and hopanes. Thereby, a clear distinction could be made between the input from the oil deliberately discharged, and that arising from natural biological reactions or mediated by atmospheric transport. 2. Ice scouring of the intertidal zone in arctic waters makes this virtually sterile. This was noted more than 170 years ago by Keilhau (1831)—so that attention was directed to components of the subtidal zone to which little attention had previously been directed, and which was expected to be particularly sensitive to oil spills. Changes in the components of the macrobenthos including infauna, epibenthos, and macroalgae were examined, and attention was also directed to the histopathological and biochemical responses of bivalve molluscs that were affected in different ways by exposure to the dispersed and the undispersed oil.

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The original reports should be consulted for details of the conclusions drawn from this comprehensive investigation, but the following seem of particular importance for this specific class of pollutant. The study provided: • • • •

A basis for determining remedial action in the event of oil spill An evaluation of the advantage or disadvantage of using oil dispersants An estimate of the persistence of oil washed ashore An assessment of the effectiveness of procedures for removing stranded oil from beaches

Experience with some major spills of crude oil has been summarized (Atlas 1995; Swannell et al. 1996), and attention was drawn to a number of important general issues. 1. Rates of biodegradation are generally low compared with the speed and ease of mechanical removal, and are strongly dependent on whether the beaches are sandy or rocky. 2. The form in which nutrients are administered, for example, granulated or slow release is important, and a periodic monitoring program should be implemented to determine the need for addition of nutrients. This is important in view of the possible adverse effects resulting from excess addition of nutrients, for example, abnormally extensive algal growth and high bacterial activity leading to oxygen limitation. 3. It is important to be able to assess biodegradation of the oil relative to that of its most recalcitrant components. A review (Sugai et al. 1997) of the remediation following the oil spill from the Exxon Valdez that includes a retrospective analysis revealed a change in the populations of alkane-degrading and aromatic hydrocarbon-degrading bacteria that could be attributed to the enrichment effect of oleophilic fertilizers and terrestrial alkane waxes. However, the long-term effects are still being debated (Renner 2006). It is worth drawing attention to marine bacteria of the genus Marinobacter (Gauthier et al. 1992) that is able to degrade alkanes and strains of Cycloclasticus that are able to degrade aromatic hydrocarbons. These are discussed in Chapter 4.

CONCLUSION There seems to be little doubt that the biodegradation of many components of petroleum can be facilitated by addition of inorganic nutrients—primarily nitrogen—and provision of suitable oxidant(s). It seems that inoculation with exogenous organisms, even when these have been isolated from the same site, is not generally effective, and that indigenous organisms are effective provided that suitable nutrients and a supply of oxygen (as O2 or H2O2) are available. Greatest effectiveness would probably be achieved by optimizing the capacity of these indigenous hydrocarbon-degrading organisms that are virtually ubiquitous.

REFERENCES Atlas RM (1981) Microbial degradation of petroleum hydrocarbons: an environmental perspective Microbiol Revs 45: 180–209. Atlas RM (1995) Bioremediation of petroleum pollutants. Int Biodet Biodeg 35: 317–327. Bossert ID, GC Compeau (1995) Cleanup of petroleum hydrocarbon contamination in soil. In Microbial transformation and Degradation of Toxic Organic Chemicals (Eds LY Young and CE Cerniglia), pp. 77–125. Wiley-Liss, New York. Braddock JF, ML Ruth, PH Catterall, JL Walworth, KA McCarthy (1997) Enhancement and inhibition of microbial activity in hydrocarbon-contaminated arctic soils: implications for nutrient-amended bioremediation. Environ Sci Technol 31: 2078–2084. Britton LN (1984) Microbial degradation of aliphatic hydrocarbons. In Microbial Degradation of Organic Compounds (Ed DT Gibson), pp. 89–129. Marcel Dekker Inc., New York.

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Cassidy DP, RL Irvine (1997) Biological treatment of a soil contaminated with diesel fuel using periodically operated slurry and solid phase reactors. Water Sci Technol 35(1): 185–192. Clemente JS, MD Mackinnon, PM Fedorak (2004) Aerobic biodegradation of two commercial naphthanic acids preparations. Environ Sci Technol 38: 1009–1016. Deschênes L, P Lafrance, J-P Villeneuve, R Samson (1996) Adding sodium dodecyl sulfate and Pseudomonas aeruginosa UG2 biosurfactants inhibits polycyclic hydrocarbon biodegradation in a weathered creosote-contaminated soil. Appl Microbiol Biotechnol 46: 638–646. Déziel É, G Paquette, R Villemur, F Lépine, F, J-G Bisaillon (1996) Biosurfactant production by a soil Pseudomonas strain growing on polycyclic aromatic hydrocarbons. Appl Environ Microbiol 62: 1908–1912. Dott W, D Feidieker, M Steiof, PM Beckerm, P Kämpfer (1995) Comparison of ex situ and in situ techniques for bioremediation of hydrocarbon-polluted soils. Int Biodet Biodeg 35: 301–316. Gauthier MJ, B Lafay, R Christen, L Fernandez, M Acquaviva, P Bonin, J-C Bertrand (1992) Marinobacter hydrocarbonoclasticus gen. nov., sp. nov., a new extremely halotolerant, hydrocarbon-degrading marine bacterium. Int J Syst Bacteriol 42: 568–576. Hutchins SR, DE Miller, A Thomas (1998) Combined laboratory/field study on the use of nitrate for in situ bioremediation of a fuel-contaminated aquifer. Environ Sci Technol 32: 1832–1840. Keilhau BM (1831) Reise i Öst- og Vest-Finnmarken samt til Beeren-Eiland og Spitsbergen i Aarene 1827 og 1828. Christiana, Norway. Margesin R, F Schinner (1997) Efficiency of indigenous and inoculated cold-adapted soil microorganisms for biodegradation of diesel oil in alpine oils. Appl Environ Microbiol 63: 2660–2664. Morgan P, RJ Watkinson (1994) Biodegradation of components of petroleum. In Biochemistry of Microbial Degradation, pp. 1–31. Kluwer Academic Publishers, Dordrecht. Mulroy PT, L-T Ou (1998) Degradation of tetraethyllead during the degradation of leaded gasoline hydrocarbons in soil. Environ Toxicol Chem 17: 777–782. Owens EH, JR Harper, W Robson, PD Boehm (1987) Fate and persistence of crude oil stranded on a sheltered beach. Arctic 40(1): 109–123. Pirnik MP (1977) Microbial oxidation of methyl branched alkanes. Crit Rev Microbiol 5: 413–422. Renner R (2006) Exxon Valdez oil no longer a threat? Environ Sci Technol 40: 6188–6189. Rontani J-F, MJ Gilewicz, VD Micgotey, TL Zheng, PC Bonin, J-C Bertrand (1997) Aerobic and anaerobic metabolism of 6,10,14-trimethylpentadecan-2-one by a denitrifying bacterium isolated from marine sediments. Appl Environ Microbiol 63: 636–643. Roussis SG, JW Fedora (1997) Determination of alkenes in hydrocarbon matrices by acetone chemical ionization mass spectrometry. Anal Chem 69: 1550–1556. Scott AA, MD Mackinnon, PM Fedorak (2005) Naphthenic acids in Athabasca oil sands tailings water are less biodegradable than commercial naphthenic acids. Environ Sci Technol 39: 8388–8394. Sergy GA (Ed) (1987) The Baffin Island Oil Spill (BIOS) Project. Arctic 40(1): 1–279. Shen Y, LG Stehmeier, G Voordouw (1998) Identification of hydrocarbon-degrading bacteria in soil by reverse sample gene probing. Appl Environ Microbiol 64: 637–645. Stehmeier LG, TR Jack, G Voordouw (1996) In vitro degradation of dicyclopentadiene by microbial consortia isolated from hydrocarbon-contaminated soil. Can J Microbiol 42: 1051–1060. Sugai SF, JE Limndstrom, JF Braddock (1997) Environmental influences on the microbial degradation of Exxon Valdez oil on the shorelines of Prince William Sound, Alaska. Environ Sci Technol 31: 1564–1572. Swannell RPJ, K Lee, M McDonagh (1996) Field evaluations of marine oil spill bioremediation. Microbiol Rev 60: 342–365. van der Linden AC (1978) Degradation of oil in the marine environment. In Development in Biodegradation of Hydrocarbons—1 (Ed RJ Watkinson), pp. 165–200. Applied Science Publishers Ltd, London.

PART 2 COAL DISTILLATION PRODUCTS INTRODUCTION Coal gas for illumination—and later for heating—was produced by the destructive distillation of coal and resulted in the production of large volumes of tar. This is used in the form of creosote for wood preservation, so that both gasworks sites and impregnation facilities became heavily contaminated with coal tar products. The range of compounds involved is enormous, but attention

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has been directed particularly to groups with established toxicity or persistence: PAHs, azaarenes, thiaarenes, and phenolic compounds—mainly cresols, xylenols, and 2,3,5-trimethylphenol (Bayly et al. 1988; Mueller et al. 1989). It should be emphasized that as a result of analytical problems due to lack of authentic samples (Biggs and Fetzer 1996), the frequency and concentrations of PAHs with more than seven rings have probably been seriously underestimated. The principal reasons for attempting remediation of such sites is the established human carcinogenicity of important PAHs including benzo[a]pyrene, dibenz[a,h]anthracene, and the dibenzopyrenes — dibenzo[a,h]pyrene, dibenzo[a,i]pyrene, and dibenzo[a,l]pyrene. PAHs with five and six rings should therefore be the primary target in bioremediation programs. It is important to emphasize that, even under favorable conditions, complete removal of PAHs—particularly those with more than five rings will seldom be accomplished: it is therefore important to predetermine an acceptable level of removal. Although the main thrust of this chapter is directed to biological processes, it should be noted that subcritical water at 275°C was successful in removing PAHs with two to six rings from a contaminated site (including the carcinogenic benzo[a]pyrene), and that the wastewater was considered suitable for further exploitation (Lagadec et al. 2000). Several reviews that cover various important aspects of the bioremediation of PAH-contaminated sites have been devoted to this topic. All of them provide valuable summaries of the various technologies that have been exploited, and cover important engineering and geological aspects including the following: a. A discussion of the design of different systems and physical factors such as temperature, pH, nutrients, and oxygen supply (Thomas and Lester 1993). b. A summary of the maximum concentrations that must be attained at contaminated sites after bioremediation, and discussion and exemplification of the various strategies that may be used—in situ, on-site, and bioreactors (Wilson and Jones 1993). c. An evaluation of a range of available strategies for remediation (Ram et al. 1993). d. A discussion of factors determining the effectiveness of bioremediation including the effect of complex mixtures of contaminants, the limitation of bioavailability imposed by aging, the value of surfactants, and of nutrients and inoculation (Hughes et al. 1997). e. A desirable procedure using extrapolation of data from laboratory experiments has been given (Findlay et al. 1995). The mobility of both bacteria and their substrates at a coal tar–contaminated site has been described (Madsen et al. 1996a), and involved the ingenious use of sorbents to trap both the bacteria and the analytes. It also drew attention to the important issue of the extent to which microorganisms penetrate the contaminated matrix. The extensive background on the biodegradation and biotransformation of PAHs has been reviewed in Chapter 8, Part 2, and should be consulted for more detailed metabolic and enzymatic aspects. The regulation of metabolism has already been noted, and the introduction of salicylate as inducer for the degradation of naphthalene (Ogunseitan and Olson 1993) has been examined. Although cells of Pseudomonas saccharophila P15 could not grow with fluoranthene, pyrene, benz[a]anthracene, chrysene, or benzo[a]pyrene, the organism was able to metabolize them, and showed considerably enhanced rates of removal and mineralization after induction with salicylate (Chen and Aitken 1999). An attempt is made to illustrate some of the complex factors that influence the degree of success of bioremediation programs.

ON-SITE OR OFF-SITE TREATMENT On-site treatment in batch reactors using refinery and wood-preserving waste added to soil resulted in highly variable loss of PAHs over a year (Aprill et al. 1990). An extensive study (Ellis et al. 1991) examined a number of design features, and the relative merits of on-site and off-site treatments of a

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creosote-contaminated site. Preliminary experiments were carried out in liquid culture and in soil pots, and showed only limited advantage of adding strains isolated from the site, and rather variable success in removal of most PAHs except naphthalene. Two kinds of large-scale experiments were also carried out a. On-site by recirculation of groundwater, and addition of surfactants, hydrogen peroxide, and inorganic nutrients b. Off-site in concrete-lined basins with recirculation of water and mixing by surface ploughing There were differences between on-site and off-site removal, and for most PAHs studied the degree of removal was variable for PAHs with up to four rings (Table 14.1). Data for the important group of PAHs with more than four rings were unfortunately not reported. A system that combined biological and chemical treatment for destruction of PAHs in contaminated soil has been described (Pradhan et al. 1997). This took advantage of the activity of the indigenous microflora in the contaminated soil in bioreactors, with an intermediate chemical reactor that used a mixture of ferrous sulfate and hydrogen peroxide (Fenton’s reagent) to carry out oxidation to more readily degradable substrates. The efficiency in removal of a range of PAHs by this process compared with conventional biological treatment is clearly shown by the results in Table 14.2.

TABLE 14.1 Comparison of the Efficiency of In-Situ and Off-Site Procedures for Removal of Selected PAHs Mean Percent Removal Substrate Naphthalene Acenaphthalene Phenanthrene Anthracene Fluoranthene Pyrene Chrysene

On-Site

Off-Site

100 3 8 36 53 37 76

100 87 74 69 57 53 50

TABLE 14.2 Comparison of the Removal of PAHs from Contaminated Soil by Conventional Biological Treatment, or by Combined Biological and Chemical Treatment PAH Number of Rings Two Three Four Five Six

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Conventional Treatment

Combined Treatment

31 76 69 22 0

96 97 95 87 88

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THE PHYSICAL STATE OF THE CONTAMINANT At many sites, the contaminants may have lain undisturbed for many years and, during this time significant alterations may have taken place. These processes include not only degradation, but also loss of the more volatile components such as naphthalene and alkylated naphthalenes, and association with the surrounding matrix. A number of investigations have examined the effect of aging on the biodegradability of PAHs. The bioavailability of xenobiotics and its decrease with aging have been considered as limiting factors in their biodegradation (Hatzinger and Alexander 1995), and this is discussed below in the context of the role of surfactants. The results of Kan et al. (1998) provide a graphic illustration of the differences in desorption predicted from a reversible and an irreversible model, and the lack of correlation between chemical extractability and bioavailability (Chung and Alexander 1998). Further aspects of this are included in an evaluation of the application of surfactants. 1. Laboratory studies on the rate of degradation of phenanthrene using samples from a contaminated site showed that this depended critically on the source of the inoculum within the site (Sandoli et al. 1996). 2. Most contaminated sites are highly heterogeneous in the concentration of the contaminant. It has been shown that for naphthalene contained in coal tar globules, the area-dependent mass transfer coefficient for globules was 103 greater than when the substrate was coated on microporous silica beads, and that this was an important factor in determining the rate of mineralization (Ghoshal and Luthy 1996). 3. The sorption–desorption of PAHs has been extensively investigated, and the role of desorption in determining their biodegradability in aged sediments has been widely accepted (references in Carmichael et al. 1997). A definitive study using 14C-phenanthrene and 14 C-chrysene showed that, in contaminated soils, their rates of mineralization were much lower than the rates of desorption from spiked sediments. By contrast, for aged substrates, desorption rates were essentially comparable to rates of mineralization. This suggested that the indigenous microflora might have become adapted to the low substrate concentrations available by desorption (Carmichael et al. 1997). This would be consistent with the existence of bacteria capable of using low substrate concentrations, but would imply the limited effectiveness of adding bacteria which—although metabolically active—were not necessarily adapted to the use of low substrate concentrations. 4. Several important issues emerged from the results of experiments with 14C-labeled pyrene added to a pristine forest soil (Guthrie and Pfaender 1998): a. Extensive mineralization took place only in samples amended with a pyrene-degrading microbial community. b. Compared with an azide-treated control, there was a substantially greater nonextractable fraction of label in soils containing either the natural or an introduced microflora. This illustrates the role of microorganisms in bringing about association. c. Metabolites that could be released by acid and base extraction were persistent in the soil even after 270 d incubation. 5. The remediation of PAHs at a former gasworks site was examined using fungi that were isolated from the site. The most effective were strains of Coniothyrium sp. and Fusarium sp. Incubation with mycelia was more effective than with spores, and this would be consistent with the significance of mycelia in penetrating the matrix. One interesting feature was the loss of PAHs with five and six rings, including the carcinogenic benzo[a]pyrene (Potin et al. 2004). These results clearly indicate the existence of a number of issues that must be resolved in the application of bioremediation to PAH-contaminated soils, and that caution should be exercised in making generalizations.

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ENVIRONMENTAL PARAMETERS Two aspects of general significance have been examined: the water concentration and the oxygen concentration. Extensive studies using microcosm experiments with creosote-contaminated soil revealed losses in surface soil samples that exceeded those from the sediment, and that the use of soil slurries was more effective than solid-phase systems. The removal of benz[a]anthracene, benzo[b/k]fluoranthene, and indeno[1,2,3-c,d]pyrene was poor even using the slurries, although the heteroarenes— quinolines, acridine, and carbazole—were partially removed (Mueller et al. 1991a,b). Laboratory experiments using spiked soil indicated that biodegradation by inoculation with bacterial strains that could degrade anthracene and pyrene was inhibited by the introduction of a mineral medium, and occurred effectively only after the addition of water (Kästner et al. 1998). These observations are in general agreement with the results of a study with Pseudomonas putida mt-2 during degradation of toluene (Holden et al. 1997). Matric water potential results from the interaction of soil water with solid surfaces, and is generally the major component of total water potential: growth rates showed a shallow maximum at a matric water potential of −0.25 MPa thereafter decreasing steadily to −1.5 MPa. Whereas oxygen must be accessible, it has been shown that the degradation of pyrene by a strain of Mycobacterium sp. can occur at low oxygen concentration (Fritzsche 1994). This is consistent with the observation that toluene-degrading bacteria isolated from sites such as aquifer sand or groundwater with low concentrations of oxygen, have lower values of Km for O2 and for 3-methylcatechol than do other strains (Kukor and Olsen 1996). During bioremediation of creosote-contaminated soil, it was shown that aeration was an important factor (Vinas et al. 2005). The biodegradation of naphthalene in a subsurface site that was oxygen-limited was not facilitated by addition of nitrate as alternative electron acceptor (Madsen et al. 1996b). In contrast, degradation of anthracene, phenanthrene, and pyrene has been shown using three strains of denitrifying pseudomonads under both aerobic and denitrifying conditions (McNally et al. 1998). This is in contrast to the effectiveness of nitrate as an electron acceptor for the degradation of monocyclic aromatic hydrocarbons in groundwater, which is noted in a later section. It is important to note that, under oxygen limitation, cultures of even strictly anaerobic methanogenic bacteria may coexist with aerobic bacteria (Gerritse and Gottschal 1993). An important observation was the toxicity of naphthalene to the naphthalenedegrading strain of Pseudomonas putida G7 under conditions of oxygen or nitrogen limitation, since both of these conditions frequently prevail at contaminated sites (Ahn et al. 1998). The physical design of any system therefore requires careful consideration, and aeration may be expensive, but obligatory.

DETERMINANTS The Range of Substrates It should be emphasized that the development of suitable chromatographic columns for gas chromatography (GC) analysis has begun to reveal the presence of PAHs with molecular masses >300. Those with more than seven rings are probably ubiquitous, although their determination is made difficult by the inaccessibility of reference compounds and by their long retention times on GC columns (Biggs and Fetzer 1996). PAHs with molecular mass >300 that occur in both carbon black and the NIST SRM 1597 coal tar extract include dibenzo[a,j]coronene (Mr 400) with nine rings, dinaphtho[2,1-a: 2,1-h]anthracene (Mr 378), and benzo[a]coronene (Mr 350) with eight rings (Bemgård et al. 1993). The degradability of such compounds is unknown, and they may possibly contain carcinogenic representatives. Attention should be directed to those PAHs that have been accepted as human carcinogens after metabolic activation: these include benzo[a]pyrene and dibenz[a,h]anthracene. In general, degradation of three- and four-ring PAHs is readily accomplished, whereas those with five or more rings may be recalcitrant. However, a study using endogenous

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bacteria in a creosote-contaminated soil showed that benz[a]anthracene and chrysene were substantially degraded, although the degradation of chrysene was less when nutrients or a biosurfactant were added (Vinas et al. 2005). It is important to emphasize that not only PAHs may be present, but also the corresponding analogs containing nitrogen (azaarenes) and sulfur (thiaarenes) in the rings. This important group of contaminants has been reviewed in detail by Herod (1998), and some azaarenes such as dibenz[a,h]acridine and 7H-dibenzo[c,g]carbazole are carcinogenic. Although the biodegradation of these has been presented in Chapter 10 and in a review (Schwarz and Lingens 1994), the possibility of remediating sites contaminated with heteroarenes presents a number of unresolved issues. The fate of azaarenes seems not to have been examined directly in this context, while that of thiaarenes has attracted attention mainly in connection with the desulfurization of coal and coal oil (Wang et al. 1996; Macpherson et al. 1998). Azaarenes do, however, have an inhibitory effect on the biodegradation of PAHs even by a versatile strain such as Sphingomonas paucimobilis strain EPA505 (Lantz et al. 1997). By comparison with their carbocyclic analogs, rather less is known about the biodegradation of azaarenes, although the biochemistry of the aerobic degradation of quinolines (including methyl- and carboxyquinolines) has been extensively investigated by Lingens and his colleagues (Schwartz and Lingens 1994). As a broad generalization, it may be stated that the initial reaction frequently involves formal hydroxylation (addition of the elements of H2O followed by dehydrogenation) in the hetero ring, whereas fission of the carbocyclic ring is accomplished by dioxygenation. It is worth noting that these organisms may be quite restricted in their ability to use other azaarenes. For example, a strain of Rhodococcus sp. that was isolated by enrichment of pristine samples with quinoline was unable to use isoquinoline or acridine, and supported only poor growth with a series of pyridines (O’Loughlin et al. 1996). Although the degradation of acridine and phenanthridine seems to have been examined only parenthetically, the degradation of carbazole has been examined more extensively. An exhaustive study has revealed the production of several terminal metabolites formed by cyclization of the fission products from the carbocyclic ring (Gieg et al. 1996). However, benzocarbazoles seem not to have been examined. The mechanism of the degradation of dibenzofurans has been established in detail and is accomplished by angular dioxygenation with fission of the furan ring, while the biodegradation of benzothiophenes is complicated by oxidation at the sulfur atom. The biotransformation of dimethylbenzothiophenes was studied in three strains of Pseudomonas sp. growing with 1-methylnaphthalene or glucose (Kropp et al. 1997). A number of oxidations of different isomers was observed: (a) methyl groups to groups and carboxylic acids, (b) at the sulfur atom to produce sulfoxides and sulfones, and (c) the thiophene ring to form the 2- and 3-hydroxy compounds [2(3H) ones, 3(2H) ones], and 2,3-diones. The same pseudomonads were able to degrade 3,4-dimethyl- and 2,7-dimethyldibenzothiophenes by oxidative fission of one of the benzene rings, followed by oxidation of the thiophene ring. The Range of Degradative Bacteria Bacteria vary considerably in their capacity to degrade a range of PAHs, and extrapolation of results from the more-readily degraded naphthalene and phenanthrene should be carried out with caution. Some bacteria degrade only the more readily degradable naphthalene and phenanthrene, although pyrene and fluoranthene seem to be more generally degradable. For example, Pseudomonas paucimobilis strain EPA 505 can degrade fluoranthene, pyrene, benz[a]anthracene, chrysene, benzo[a]pyrene, and benzo[b]fluoranthene, (Ye et al. 1996), and Gram-positive mycobacteria have attracted increasing attention: for example, a strain of Mycobacterium sp. degraded pyrene, benz[a]anthracene, and benzo[a]pyrene (Schneider et al. 1996). Details of these are given in Chapter 8, Part 2. In practice, contaminated sites will maintain a wide range of bacteria, and it has been shown that the flora will alter during degradation and will be influenced by the addition of nitrogen and phosphorus nutrients (Vinas et al. 2005). In that study, denaturing gradient gel electrophoresis (DGGE) revealed a range of taxa. These included organisms belonging to the α-Proteobacteria

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(Sphingomonas) at the early stages of biodegradation, whereas at later stages in the absence of nitrogen and phosphorus supplementation these were supplemented by members of the γ-Proteobacteria (Xanthomonas) and the Cytophaga–Flexibacter–Bacteroides group (Bacteroides). Bioavailability and the Use of Surfactants This is a cardinal issue with many, sometimes contradictory facets, that merits more detailed consideration than can be given here. Strategies that have been used by cells for substrates with low or negligible water solubility are relevant to strategies for bioremediation. The accessibility of substrates at sites with a long history of contamination is a serious limitation to bioremediation, and attempts have been made to increase this by the addition of surfactants (Grimberg et al. 1996). A review (Volkering et al. 1998) has summarized data from experiments with alkanes and PAHs that used a range of surfactants. A few of the sometimes-contradictory aspects of the role of surfactants are briefly summarized: 1. Significant differences have been found among different organisms (Guerin and Boyd 1992b; Crocker et al. 1995), and indeed incorporation of PAHs into micelles may make them less accessible as substrates (Volkering et al. 1995). 2. It has been shown in biodegradation experiments using a mixed bacterial culture, that components in a mixture of naphthalene, phenanthrene, and pyrene were bioavailable from the micellar phase of Triton X-100, although the degree of bioavailability decreased with increasing surfactant concentration (Guha et al. 1998). A model was constructed that could be used to determine optimal surfactant concentration for criteria including maximum rates of biodegradation. 3. There is evidence that, as for alkanes, biosurfactants may be produced during growth with PAHs (Déziel et al. 1996), although the practical implementation of this may be limited by the requirement of a suitable monomer such as glucose or mannitol for their synthesis. Pseudomomas aeruginosa GL1 that was part of a community that degraded PAHs, produced a rhamnolipid in the late stage of growth with glycerol or n-hexadecane and inhibited the growth of Bacillus cereus and Rhodococcus erythropolis (Arino et al. 1998). 4. The biodegradation of pyrene, chrysene, fluoranthene, benz[a]anthracene, dibenz[a,h] anthracene, benzo[a]pyrene, and coronene by Stenotrophomonas maltophilia has been studied in the presence of a range of synthetic surfactants (Boonchan et al. 1998). Nonneutral surfactants were toxic, biodegradation was also inhibited by the neutral Igepal CA-630, and the positive enhancement of removal of substrates was generally low—in the range of 10%. 5. A comparison was made of the mineralization of 14C-fluoranthene by two strains each of Mycobacterium sp. and Sphingomonas sp. (Willumsen et al. 1998). Triton X-100 decreased the rate of mineralization by the Mycobacterium strains, inhibited mineralization by Sphingomonas sp. strain EPA505, and diminished the survival of both Gram-negative strains during 50 h incubation. It is worth noting that the Sphingomonas strain EPA 505 is able to degrade a wide range of PAHs (Ye et al. 1996), and would therefore otherwise be a potential candidate for the bioremediation of PAH-contaminated sites. The differential effects of the neutral surfactants 2,2,4,4,6,8,8-heptamethylnonane and Triton X-100 on the formation of cis-dihydrodiols of naphthalene and phenanthrene by Pseudomonas sp. strain 9816/11 and Sphingomonas yanoikuyae strain B8/36 further illustrated the significance of both the organism and the structure of the surfactants (Allen et al. 1999). 6. During degradation of artificial surfactants, the consumption of oxygen may, however, result in diminished PAH degradation (Tiehm et al. 1997), so that surfactants with lower biodegradability would be advantageous from this point of view.

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7. A lipoprotein surfactant is produced during growth of Pseudomonas marginalis strain PD-14B with succinate, and prevented flocculation of cells in media containing anthracene, acenaphthylene, naphthalene, and chrysene (Burd and Ward 1996). 8. Increased removal of phenanthrene from soil columns spiked with the rhamnolipid mixture synthesized by Pseudomonas aeruginosa UG2 has been demonstrated, and shown to depend both on the increased desorption of the substrate and on partitioning into micelles (Noordman et al. 1998). However, the addition of the biosurfactant from the same strain of Pseudomonas aeruginosa UG2 or of sodium dodecyl sulfate had no effect on the rate of biodegradation of anthracene and phenanthrene from a chronically contaminated soil. At concentrations of sodium dodecyl sulfate >100 mg/kg, there was a marked inhibition of the degradation of fluoranthene, pyrene, benz[a]anthracene, and chrysene (Deschênes et al. 1996). It was suggested that this was due to competition with the increasingly dominant surfactant-degrading microorganisms. This conclusion is supported by evidence that the addition of nonionic surfactants (alcohol ethoxylates) was more effective in enhancing biodegradation of PAHs in coal tar–contaminated soil than the more readily degradable glycoside surfactants (Madsen and Kristensen 1997). 9. The addition of a rhamnolipid biosurfactant produced by Pseudomonas aeruginosa stain AT10 apparently reduced the extent of degradation by endogenous bacteria of benz[a]anthracene and chrysene in a creosote-contaminated soil (Vinas et al. 2005). The relation between the type and concentration of surfactant and the stimulation of biodegradation is therefore complex, and has not so far been fully resolved (Burd and Ward 1996). There seems to be a consensus, however, that supplementation with surfactants is one of the factors that might enhance bioremediation of some contaminated sites (Walter et al. 1997). Some cardinal issues that are potentially negative include the following (Willumsen et al. 1998): • Diminished bioavailability of substrates from micelles • Toxicity or inhibition of biodegradation by some nonneutral synthetic surfactants as a result of surfactant-induced permeabilization of the cells • Toxicity as a result of the surfactant-enhanced PAH concentration in the aqueous phase • Adverse effect of surfactants on bacterial adhesion to surfaces in the matrix • Adverse effect of readily degraded surfactants that diminish the oxygen concentration available to the substrate-degrading organisms • The need to maintain a high density of the substrate-degrading organisms

SUPPLEMENTATION WITH FUNGI Some preliminary comments are given to provide perspective: a. It is important to emphasize that, in contrast to bacterial degradation of PAHs, many fungi carry out only transformation without ring fission. An important consequence is that demonstration of loss during bioremediation experiments is equivocal. A monooxygenase produces an epoxide that is either hydrolyzed to a trans-dihydrodiol or rearranged to a phenol that is often conjugated. It has been suggested that naphth-1-ol, which is one of the expected transformation products of naphthalene is the form in which at least part of naphthalene is associated with soil matrices (Burgos et al. 1996). b. The hydrogen peroxide that is required by white-rot fungi is produced by a range of reactions that include the activity of several oxidases and, in Ceriporiopsis subvermispora by manganese peroxide–dependent oxidation of glycolate and oxalate that are synthesized by the organism (Urzúa et al. 1998).

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c. Although considerable attention has been directed to the application of white-rot fungi, several important issues should be noted. Most attention has hitherto been given to species of Phanerochaete especially Ph. chrysosporium and Ph. sordida, although other species and other taxa (Field et al. 1992) may be of greater interest in view of their greater biodegradative capability. d. Whereas transformation has been demonstrated—often to quinones—only relatively low levels of mineralization to CO2 have frequently been observed, for example, in Pleurotus ostreatus (Bezalel et al. 1996a,b; Hofrichter et al. 1998). Illustrative examples of the application of fungi include the following: 1. The use of fungi has already been noted (Potin et al. 2004), although no analysis of putative metabolites was carried out. This is important since there has been considerable interest in the use of fungi in bioremediation programs, in which quinones may be produced as transient or terminal metabolites. For example, although anthracene-9,10-quinone is produced by Pleurotus ostreatus strain MUCL 29257 as a terminal metabolite from anthracene, it is only transiently formed by Phanerochaete chrysosporium strain DSM 1556 (Andersson and Henrysson 1996). In a study with spiked soil, transformation by Pleurotus ostreatus produced fluorenone, benz[a]anthracene-7,12-quinone, and 4-hydroxyperinaphthenone (Andersson et al. 2003). White-rot fungi belonging to other taxa may, however, be more effective in mineralization. For example, Nematoloma frowardii mineralized 8.6% of the 14C-labeled phenanthrene supplied, although a substantial portion (46%) was associated with the mycelia (Sack et al. 1997; Hofrichter et al. 1998). A few examples of bioremediation using fungi are given as illustration. The supplementation of indigenous bacteria with Phanerochaete chrysosporium increased the loss of phenanthrene at a former oil gasification site, although there was apparently a significant contribution to “polar metabolites” (Brodkorb and Legge 1992). Phenanthrene is not, however, representative of the toxicological profile of PAHs, although the presence of unidentified metabolites may plausibly be interpreted as evidence of transformation. Application of the white-rot fungus Phanerochaete sordida to treat creosote-contaminated soil has been described, but again had limited success with PAHs having more than four rings (Davis et al. 1993). 2. Not only white-rot fungi have been examined, but comprehensive studies also have been made using Cunninghamella sp. (references in Sutherland et al. 1995). Experiments in microcosms supplemented with sources of inorganic nitrogen and phosphorus using the nonlignolytic fungus Cunninghamella echinulata var. elegans showed not only the expected loss of three- and four-ring PAHs but also partial loss of benzo[b]fluoranthene and benzo[g,h,i]perylene (Cutright 1995). In this study, hydrogen peroxide was used as source of oxygen, although in natural systems the presence of bacteria with catalase activity would probably effectively destroy the hydrogen peroxide. 3. An important series of observations with a PAH-contaminated soil (May et al. 1997) showed that, although the fraction that was mineralized was low (ca. 2.5%), the PAHs were lost by polymerization processes mediated by the fungus. It was suggested that this was consistent with (a) the formation of pyrene-1,6-quinone and pyrene-1,8-quinone from pyrene (Hammel et al. 1986) and the 1,6-, 3,6-, and 6,12-quinones from benzo[a]pyrene by this organism (Haemmerli et al. 1986); (b) the tendency of quinones to polymerize and associate with soil humic material. The loss of the higher PAHs was highly variable, ranging from ca. 70% for benzo[b]fluoranthene to ca. 5.5% for benzo[a]pyrene. 4. It is worth drawing attention to the significance of other issues. In natural ecosystems, other microorganisms including bacteria are almost always present and it has been shown that, in experiments using [7,10-14C]benzo[a]pyrene, incubation for 215 d with Bjerkandera sp. strain BOS55 alone resulted in the formation of 13.5% 14CO2 and 61% of labeled metabolites in the

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water phase. Addition of microbial cultures from forest soil at day 15 increased the 14CO2 evolution to 32% and reduced the water-soluble metabolites to 18% (Kotterman et al. 1998).

APPLICATION OF HIGHER PLANTS 1. The use of hybrid poplar (Populus deltoides × P. nigra) has been examined in a few investigations at sites contaminated with creosote. There are several issues that complicate the interpretation of these results, including the physical state of the contaminants, the mechanism of uptake from the soil, and translocation and metabolism in the plants. These have seldom been addressed in studies of PAHs. a. A push–pull study illustrated differences in respiration between sites that had been planted and those that had not, although data were not presented in terms of specific PAH components (Pitterle et al. 2005). b. A 7-year study established a diminution in the PAH plume, although the major decline in specific PAHs was limited primarily to naphthalene, and the occurrence of a dense liquid phase of creosote (Widdowson et al. 2005). 2. A site polluted for over 20 years with hydrocarbons was assessed for improvement using a range of plants, including grasses and legumes, some of them native species. The analysis combined mineralization of 13C-labeled naphthalene in the bulk soil and in the rhizosphere with the occurrence of genes specific for dioxygenation of naphthalene (ndoB), alkane monooxygenation (alkB), and catechol 2,3-dioxygenase (xylE). Festuca arundinacea (Tall fescue) brought about an increase in the prevalence of the genes and mineralization of naphthalene in soil from the rhizosphere, whereas Trifolium hirtum (Rose clover) diminished both (Siciliano et al. 2003). There were several important conclusions: (a) aged contaminants were susceptible to plant-mediated degradation, (b) the results were plantspecific, and (c) they revealed the significance of the rhizosphere.

ANOXIC OR ANAEROBIC ENVIRONMENTS At many sites, the subsurface environment will be anoxic or even anaerobic due to the activity of aerobic and facultatively anaerobic bacteria in the surface layers of the soil. It is therefore essential to take into consideration the extent to which anaerobic degradation may be expected to be significant. Reactions may take place under sulfidogenic or methanogenic conditions, and the occurrence of sulfate at sites containing building material waste and the metabolic versatility of sulfate-reducing bacteria makes them particularly attractive. Hydrocarbons Hitherto, studies on the anaerobic degradation of aromatic hydrocarbons have been limited to monocyclic representatives—particularly toluene and xylenes. This issue has been discussed more fully in the context of benzene/toluene/ethylbenzene/xylene (BTEX) remediation in groundwater and the anaerobic degradation of monocyclic aromatic hydrocarbons. There is, however, evidence for the degradation of naphthalene and phenanthrene coupled to sulfate reduction under anaerobic conditions (Coates et al. 1996), and of naphthalene, phenanthrene, and pyrene under denitrifying conditions (McNally et al. 1998). It has been shown that the transformation of 2-methylnaphthalene and the cometabolism of benzothiophene and benzofuran are carried out under sulfate-reducing conditions with the formation of the polar carboxylates (Safinowski et al. 2006). These could therefore be used as markers for the occurrence of partial degradation in groundwater under anaerobic conditions in the presence of sulfate. Phenols The anaerobic degradation of phenol proceeds by carboxylation of phenyl phosphate, followed by dehydroxylation, and fission of the ring after partial reduction (Brackmann and Fuchs 1993).

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Degradation of o- (Bisaillon et al. 1991) and m-cresol (Roberts et al. 1990) has been observed under methanogenic conditions, and a summary of the possibilities under denitrifying conditions has been given (Rudolphi et al. 1991). An alternative pathway involving oxidation of the methyl groups under denitrifying conditions has also emerged (Bonting et al. 1995). Heteroarenes Although the anaerobic degradation of azaarenes containing several nitrogen atoms—particularly pyrimidines and purines—has been extensively studied, less is known about those with only a single heteroatom. Evidence for the loss of pyridine in sulfate-reducing and methanogenic aquifer slurries and of 2-methylpyridine in the former has been presented (Kuhn and Suflita 1989), although no details of the degradative pathways are available. Whereas indole and quinoline were degraded anaerobically under nitrate-reducing, sulfate-reducing, and methanogenic conditions, neither benzofuran nor benzothiophene were degraded (Licht et al. 1996).

ABIOTIC TRANSFORMATIONS Although this issue has not been frequently addressed, it is relevant to evaluating bioremediation. Several important issues emerged from a study using a range of three- and four-ring PAHs. 1. In contrast to the three-ring compounds, residues of benz[a]anthracene, chrysene, and benzo[a]pyrene were found after 15 weeks incubation in compost-amended soil. 2. Neither dihydrodiols formed by bacterial dioxygenation, nor phenols from fungal monooxygenation followed by rearrangement or hydrolysis and elimination, were found. 3. Whereas plausible fungal metabolites from anthracene, acenaphthylene, fluorene, and benz[a]anthracene—anthracene-9,10-quinone, acenaphthene-9,10-dione, fluorene-9-one, and benz[a]anthracene-7,12-quinone—were found transiently in compost-amended soil, these were formed even in sterile controls by abiotic reactions (Wischmann and Steinhart 1997). These results clearly illustrate the care that must be exercised in interpreting the occurrence of PAH oxidation products in bioremediation experiments as evidence of biological activity. All these factors should be carefully considered in the design and implementation of any bioremediation program. The rather pessimistic views presented above are supported by quoting the succinct conclusions of a study on PAH loss in laboratory microcosms using soil from a site contaminated with PAHs from a previous gas manufacturing facility (Erickson et al. 1993). a. PAHs were unavailable for microbial degradation and resistant to mineralization. b. The soils contained significant populations of bacteria that were not adversely affected by the toxicity of the contaminants. c. Addition of readily degradable naphthalene or phenanthrene did not improve degradation of the indigenous PAHs. d. PAHs were not leached into aqueous extracts from the soil. Since the indigenous microflora will have been exposed to the substrates, the relevant degradative pathways will often be present in populations of these microorganisms and induction of the appropriate degradative enzymes will have taken place. Indeed, many of the most versatile organisms have been isolated from heavily contaminated sites. If therefore, organisms with the relevant degradative capacity have been demonstrated, the prognosis for remediation is good. In contrast, although a mixed flora obtained by enrichment with creosote was able to degrade two- and three-ring PAHs, those with four or five more rings including fluoranthene that is a dominant component of creosote were unaltered (Selifonov et al. 1998). In addition, readily degradable carbon sources are probably

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counterproductive to maintaining the required degradative capacity, although addition of inorganic nutrients and provision of suitable oxygen concentrations and of water may be necessary. There are therefore several potentially serious problems that merit further critical investigation.

CONCLUSIONS There are several important conclusions that may be drawn from the results of these studies: 1. Determination of degradation by measurement of loss may be ambiguous since this fails to take into account transformation to metabolites. At field sites, loss by other processes such as volatilization is difficult to assess, but may be inferred, for example, from low concentrations of the more volatile alkylnaphthalenes. The use of stable isotopes is discussed in Chapter 13. 2. The collective evidence suggests that although PAH components with less than four rings are probably lost, those with five or more rings may be recalcitrant. 3. The presence of azaarenes—primarily quinolines, acridine, and carbazole—in creosote may inhibit or be incompatible with the degradation of PAHs. 4. The presence of high concentrations of phenols may have an adverse effect of the microflora as a result of their toxicity. 5. The extent to which bioavailability is a limiting factor and may be circumvented by addition of surfactants remains incompletely resolved. 6. The most serious limitation is probably the degree to which the substrates are accessible to the relevant microorganisms, while the advantage of adding surfactants is equivocal. 7. Oxygen limitation may be a seriously limiting factor especially in deep subsurface sites, and there is evidence that degradation in the solid phase is less effective than in slurries. 8. Supplementation with nitrogen and phosphorus may increase the total numbers of heterotropic organisms without necessarily improving degradation of PAHs. On the basis of these limitations, a cautious view of the effectiveness of bioremediation of PAHs seems justified, and they should be taken into account in the design and strategy of bioremediation.

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Gieg LM, A Otter, PM Fedorak (1996) Carbazole degradation by Pseudomonas sp LD2, metabolic characteristics and identification of some metabolites. Environ Sci Technol 30: 575–585. Grimberg SJ, WT Stringfellow, MD Aitken (1996) Quantifying the biodegradation of phenanthrene by Pseudomonas stutzeri P 16 in the presence of a nonionic surfactant. Appl Environ Microbiol 62: 2387–2392. Guerin WF, SA Boyd (1992b) Differential bioavailability of soil-sorbed naphthalene to two bacterial species. Appl Environ Microbiol 58: 1142–1152. Guha S, PR Jaffé, CA Peters (1998) Bioavailability of mixtures of PAHs partitioned into the micellar phase of a nonionic surfactant. Environ Sci Technol 32: 2317–2324. Guthrie EA, FK Pfaender (1998) Reduced pyrene bioavailability in microbially active soils. Environ Sci Technol 32: 501–508. Haemmerli SD, MSA Leisola, D Sanglard, A Fiechter (1986) Oxidation of benzo[a]pyrene by extracellular ligninases of Phanerochaete chrysosporium. J Biol Chem 261: 6900–6903. Hammel KE, B Kalyanaraman, TK Kirk (1986) Oxidation of polycyclic aromatic hydrocarbons and dibenzo[p]-dioxins by Phanerochaete chrysosporium. J Biol Chem 261: 16948–16952. Hatzinger PB, M Alexander (1995) Effect of ageing of chemicals in soil on their biodegradability and extractability. Environ Sci Technol 29: 537–545. Herod AA (1998) Azaarenes and thiaarenes. Handbook Environ Chem 3I: 271–323. Hofrichter M, K Scheibner, I Schneega, W Fritsche (1998) Enzymatic combusion of aromatic and aliphatic compounds by manganese peroxidase from Nematoloma frowardii. Appl Environ Microbiol 64: 399–404. Holden PA, LJ Halverson, MK Firestone (1997) Water stress effects on toluene biodegradation by Pseudomonas putida. Biodegradation 8: 143–151. Hughes JB, DM Beckles, SD Chandra, CH Ward (1997) Utilization of bioremediation processes for the treatment of PAH-contaminated sediments. J Ind Microbiol Biotechnol 18: 152–160. Kan AT, G Fu, M Hunter, MB Tomson (1998) Irreversible sorption of neutral hydrocarbons to sediments: experimental observations and model predictions. Environ Sci Technol 32: 892–902. Kästner M, M Breuer-Jammali, B Mahro (1998) Impact of inoculum protocols, salinity, and pH on the degradation of polycyclic aromatic hydrocarbons (PAHs) and survival of PAH-degrading bacteria introduced into soil. Appl Environ Microbiol 64: 359–362. Kotterman MJJ, EH Vis, JA Field (1998) Successive mineralization and detoxification of benzo[a]pyrene by the white-rot fungus Bjerkandera sp strain BOS55 and indigenous microflora. Appl Environ Microbiol 64: 2853–2858. Kropp KG, S Saftic, JT Andersson, PM Fedorak (1997) Transformations of six isomers of dimethylbenzothiophenes by three Pseudomonas strains. Biodegradation 7: 203–221. Kuhn EP, JM Suflita (1989) Microbial degradation of nitrogen, oxygen and sulfur heterocyclic compounds under anaerobic conditions: studies with aquifer samples. Environ Toxicol Chem 8: 1149–1158. Kukor JJ, RH Olsen (1996) Catechol 2,3-dioxygenases functional in oxygen-limited hypoxic environments. Appl Environ Microbiol 62: 1728–1740. Lagadec AJM, DJ Miller, AV Lilke, SB Hawthorne (2000) Pilot-scale subcritical water remediation of polycyclic aromatic hydrocarbon- and pesticide-contaminated soil. Environ Sci Technol 34: 1542–1548. Lantz SE, MT Montgomery, WW Schultz, PH Pritchard, BJ Spargo, JG Mueller (1997) Constituents of an organic wood preservative that inhibit the fluoranthene-degrading activity of Sphingomonas paucimobilis strain EPA505. Environ Sci Technol 31: 3573–3580. Licht D, BK Ahring, E Arvin (1996) Effects of electron acceptors, reducing agents, and toxic metabolites on anaerobic degradation of heterocyclic compounds. Biodegradation 7: 83–90. Macpherson T, CW Greer, E Zhou, AM Jones, G Wisse, PCK Lau, B Sankey, MJ Grossman, J Hawari (1998) Application of SPME/GC-MS to characterize metabolites in the biodesulfurization of organosulur model compounds in bitumen. Environ Sci Technol 32: 421–426. Madsen EL, CL Mann, SE Bilotta (1996a) Oxygen limitations and aging as explanations for the field persistence of naphthalene in coal tar-contaminated surface sediments Environ Toxicol Chem 15: 1876–1882. Madsen EL, CT Thomas, MS Wilson, RL Sandoli, SE Bilotta (1996b) In situ dynamics of aromatic hydrocarbons and bacteria capable of PAH metabolism in a coal tar-waste-contaminated field site Environ Sci Technol 30: 2412–2416. Madsen T, P Kristensen (1997) Effects of bacterial inoculation and nonionic surfactants on degradation of polycyclic aromatic hydrocarbons in soil. Environ Toxicol Chem 16: 631–637.

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May R, P Schröder, H Sandermann (1997) Ex-situ process for treating PAH-contaminated soil with Phanerochaete chrysosporium. Environ Sci Technol 31: 2626–2633. McNally DL, JR Mihelcic, DR Lueking (1998) Biodegradation of three- and four-ring polycyclic aromatic hydrocarbons under aerobic and denitrifying conditions. Environ Sci Technol 32: 2633–2639. Mueller JG, PJ Chapman, PH Pritchard (1989) Creosote-contaminated sites. Their potential for bioremediation. Environ Sci Technol 23: 1197–1201. Mueller JG, SE Lantz, BO Blattmann, PJ Chapman (1991a) Bench-scale evaluation of alternative biological treatment processes for the remediation of pentachlorophenol- and creosote-contaminated materials: solid-phase bioremediation. Environ Sci Technol 25: 1045–1055. Mueller JG, SE Lantz, BO Blattmann, PJ Chapman (1991b) Bench-scale evaluation of alternative biological treatment processes for the remediation of pentachlorophenol- and creosote-contaminated materials: slurry-phase bioremediation. Environ Sci Technol 25: 1055–1061. Noordman WH, W Ji, ML Briusseau, DB Janssen (1998) Effects of rhamnolipid biosurfactants on removal of phenanthrene from soil. Environ Sci Technol 32: 1806–1812. O’Loughlin EJ, SR Kehrmeyer, GK Sims (1996) Isolation, characterization, and substrate utilization of a quinoline-degrading bacterium. Int Biodet Biodeg 38: 107–118. Ogunseitan OA, BH Olson (1993) Effect of 2-hydroxybenzoate on the rate of naphthalene mineralization in soil. Appl Microbiol Biotechnol 38: 799–807. Pitterle MT, RG Andersen, MA Widdowson (2005) Push-pull tests to quantify in situ degradation rates at a phytoremediation site. Environ Sci Technol 39: 9317–9323. Potin O, C Rafin, E Veignie (2004) Bioremediation of an aged polycyclic aromatic hydrocarbons (PAHs)contaminated soil by filamentous fungi isolated from the soil. Int Biodet Biodeg 54: 45–52. Pradhan SP, JR Paterek, BY Liu, JR Conrad, VJ Srivastava (1997) Pilot-scale bioremediation of PAHcontaminated soils. Appl Biochem Biotechnol 63/65: 759–773. Ram NM, DH Bass, R Falotico, M Leahy (1993) A decision framework for selecting remediation technologies at hydocarbon-contaminated sites. J Soil Contam 2: 167–189. Roberts DJ, PM Fedorak, SE Hrudey (1990) CO2 incorporation and 4-hydroxy-2-methylbenzoic acid formation during anaerobic metabolism of m-cresol by a methanogenic consortium. Appl Environ Microbiol 56: 472–478. Rudolphi A, A Tschech, G Fuchs (1991) Anaerobic degradation of cresols by denitrifying bacteria. Arch Microbiol 155: 238–248. Sack M Hofrichter, W Fritsche (1997) Degradation of phenanthrene and pyrene by Nematoloma frowardii. J Basic Microbiol 4: 287–293. Safinowski M, C Griebler, RU Meckenstock (2006) Anaerobic cometabolic transformation of polycyclic and heterocyclic aromatic hydrocarbons: evidence from laboratory and field studies. Environ Sci Technol 40: 4165–4173. Sandoli RL, WC Ghiorse, EL Madsen (1996) Regulation of microbial phenanthrene mineralization in sediment samples by sorbent-sorbate contact time, inocula and gamma irradiation-induced sterilization artifacts. Environ Toxicol Chem 15: 1901–1907. Schneider J, R Grosser, K Jayasimhulu, W Xue, D Warshawsky (1996) Degradation of pyrene, benz[a]anthracene, and benzo[a]pyrene by Mycobacterium sp strain RGHII-135, isolated from a former coal gasification site. Appl Environ Microbiol 62: 13–19. Schwartz G, F Lingens (1994) Bacterial degradation of N-heterocyclic compounds. In Biochemistry of Microbial Degradation (Ed C Ratledge), pp. 459–486. Kluwer Academic Publishers, Dordrecht, The Netherlands. Selifonov SA, PJ Chapman, SB Akkerman, JE Gurst, JM Bortiatynski, MA Nanny, PG Hatcher (1998) Use of 13C nuclear magnetic resonance to assess fossil fuel biodegradation: fate of [1-13C]acenaphthene in creosote polycyclic aromatic compound mixtures degraded by bacteria. Appl Environ Microbiol 64: 1447–1453. Siciliano SD, JJ Germida, K Banks, CW Greer (2003) Changes in microbial community composition and function during a polyaromatic hydrocarbon phytoremediation field trial. Appl Environ Microbiol 69: 483–489. Sutherland JB, F Rafii, AA Kahn, CE Cerniglia (1995) Mechanisms of polycyclic aromatic hydrocarbon degradation. In Microbial transformation and degradation of toxic organic chemicals (Eds LY Young and CE Cerniglia), pp. 269–306. Wiley-Liss, New York. Thomas AO, JM Lester (1993) The microbial remediation of former gasworks sites, a review. Environ Technol 14: 1–24.

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Tiehm A, M Stieber, P Werner, FM Frimmel (1997) Surfactant-enhanced mobilization and biodegradation of polycyclic aromatic hydrocarbons in manufactured gas plant soil. Environ Sci Technol 31: 2570–2576. Urzúa U, PJ Kersten, R Vicuna (1998) Manganese peroxidase-dependent oxidation of glycolic and oxalic acids synthesized by Ceriporiopsis subvermispora produces extracellular hydrogen peroxide. Appl Environ Microbiol 64: 68–73. Vinas M, J Sabaté, MJ Espuny, AM Solanas (2005) Bacterial community dynamics and polycyclic aromatic hdrocarbon degradation during bioremediation of heavily creosote-contaminated soil. Appl Environ Microbiol 71: 7008–7018. Volkering F, AM Breure, JG van Andel, WH Rulkens (1995) Influence of nonionic surfactants on bioavailability and biodegradation of polycyclic hydrocarbons. Appl Environ Microbiol 61: 1699–1705. Volkering F, AM Breure, WH Rulkens (1998) Microbiological aspects of surfactant use for biological soil remediation. Biodegradation 8: 401–417. Walter MV, EC Neilson, G Firmstone, DG Martin, MJ Clayton, S Simpson, S Spaulding (1997) Surfactant enhances biodegradation of hydrocarbons: microcosm and field study. J Soil Contam 6: 61–77. Wang P, AE Humphrey, S Krawiec (1996) Kinetic analysis of desulfurization of dibenzothiophene by Rhodococcus erythropolis in continuous cultures. Appl Environ Microbiol 62: 3066–3068. Widdowson MA, S Shearer, RG Andersen, JT Novak (2005) Remediation of polycyclic aromatic hydrocarbon compounds in groundwater using poplar trees. Environ Sci Technol 39: 1598–1605. Widdowson MA, S Shearer, RK Andersen, JT Novak (2005) Remediation of polycyclic aromatic hydrocarbon compounds in groundwater using poplar trees. Environ Sci Technol 39: 1598–1605, Environ Sci Technol 39: 9317–9323. Willumsen PA, U Karlson, PH Pritchard (1998) Response of fluoranthene-degrading bacteria to surfactants. Appl Microbiol Biotechnol 50: 475–483. Wilson SS, KC Jones (1993) Bioremediation of soil contaminated with polynuclear aromatic hydrocarbons PAHs, a review. Environ Pollut 81: 229–249. Wischmann H, H Steinhart (1997) The formation of PAH oxidation products in soils and soil/compost mixtures. Chemosphere 35: 1681–1698. Ye D, MA Siddiq, A Maccubbin, S Kumar, HC Sikka (1996) Degradation of polynuclear aromatic hydrocarbons by Sphingomonas paucimobilis. Environ Sci Technol 30: 136–142.

PART 3 WOOD PRESERVATION SITES: CHLOROPHENOLIC COMPOUNDS Creosote contamination has already been noted, and in this section only contamination with polychlorinated phenols—particularly pentachlorophenol—will be discussed. It should, however, be noted that commercial samples of pentachlorophenol also contain lower chlorinated congeners, and possibly chlorinated dibenzodioxins dibenzofurans. Concern with the production, use, and disposal of chlorophenols is motivated by their widely occurring toxic effects (Ahlborg and Thunberg 1980) and their possible persistence. The degradation of chlorophenols has been extensively investigated by a number of workers, and the pathways have been reviewed (Häggblom 1990). The application of specific strains to remediation of a wood preservation site in Finland and elsewhere has been presented in detail (Häggblom and Valo 1995). Contamination of surrounding groundwater may also have occurred, and fluidized-bed reactors inoculated with activated sludge from pulp mill producing chemothermomechanical pulp have been used to treat groundwater highly contaminated with chlorophenols. An important aspect of this study was that the organisms were fully functional at the relatively low temperature of 5–7°C (Järvinen et al. 1994). Evaluation of white-rot fungi for treatment of pentachlorophenol-contaminated soil has been made (Lamar and Dietrich 1990; Lamar et al. 1993), and two negative aspects have emerged: the production of the highly lipophilic pentachloroanisole and the association of the substrate with the soil matrix that presents a potentially serious problem. A pilot-scale study examined the possibility of injecting low-pressure steam beneath the surface, and it was suggested that desorption of pentachlorophenol resulted in its accessibility to treatment (Tse et al. 2001). Some general microbiological issues are briefly summarized in addition to details that are given in Chapter 9, Part 2.

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AEROBIC BACTERIA The aerobic bacterial degradation of phenols with less than three chlorine atoms occurs by initial hydroxylation, followed by ring fission and elimination of chloride from the acyclic intermediates. For compounds with more than three chlorine atoms, a different mechanism has been elucidated for the degradation of pentachlorophenol by (a) Mycobacterium chlorophenolicum strain PCP1 (Apajalahti and Salkinoja-Salonen 1987a,b), (b) a strain of Flavobacterium sp. (Steiert and Crawford 1986; Xun et al. 1992), and (c) Sphingomonas chlorophenolica (Miethling and Karlson 1996). Details of the pathway can be found in a review (Häggblom 1990). In the application of such strains to bioremediation of contaminated sites, a number of important considerations should be considered. 1. The tolerance of the strains to high concentrations of pentachlorophenol—S. chlorophenolica appears to be less sensitive than M. chlorophenolicus (Miethling and Karlson 1996). This may be attributed to the ability of the cells to adapt their metabolism to avoid synthesis of toxic concentrations of chlorinated hydroquinones, and is consistent with the low levels of these metabolites measured in the cytoplasm of cells metabolizing pentachlorophenol (McCarthy et al. 1997). Inocula have also been immobilized on polyurethane that, in addition, ameliorates the toxicity of chlorophenols (Valo et al. 1990). 2. The longevity of the strains in the environment. This is dramatically illustrated from the observations that an initial inoculum of 108 cells/g of S. chlorophenolica fell to zero within 7 months, whereas 106 –107 cells/g of M. chlorophenolicum could be recovered from a similar inoculum after 7 months (Miethling and Karlson 1996). 3. There are important requirements that may be necessary, and an additional substrate may be required either to promote cell growth or to serve as a reductant. Degradation may be functional under low oxygen concentrations and be facilitated by the presence of straw compost (Laine and Jörgensen 1996). 4. Microbial O-methylation to chloroanisoles may occur, and under aerobic conditions these may be terminal metabolites. As a result of their lipophilic character they may be transported into higher biota and hence into the food chain via earthworms (Palm et al. 1991). The detection and persistence of Mycobacterium chlorophenolicum in soil that is a prerequisite for demonstrating its positive effect on biodegradation has been examined. This used 16S rRNA combined with polymerase chain reaction (PCR) (Briglia et al. 1996) that had a detection limit of ca. 3 × 102 cells/g. It should be noted that extensive investigations by Laine have shown the effectiveness of additives in facilitating the biodegradation of chlorophenols. For contaminated soils that already contain degradative microorganisms, however, the addition of an inoculum provided little or no additional advantage (Laine and Jörgensen 1997).

DEGRADATION BY FUNGI Both 2,4-dichlorophenol and 2,4,5-trichlorophenol can be degraded by the white-rot fungus Phanerochaete chrysosporium by a complex pathway that involves oxidative displacements of chloride and O-methylation with the formation of 1,2,4,5-tetrahydroxybenzene before ring fission (Valli and Gold 1991). Although pentachloroanisole is formed from pentachlorophenol by a number of species of Phanerochaete, it can subsequently be mineralized. The formation of products by oxidative coupling with loss of chloride mediated by laccase or peroxidase should also be noted (Dec and Bollag 1994). Fungal activity may also result in the formation of stable associations by reaction of xenobiotics and their metabolites with humic acids.

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ANAEROBIC DECHLORINATION Anaerobic dechlorination of halogenated phenols has been discussed in detail in Chapter 9, Part 2. It remains only to be noted that although dechlorination is a widespread reaction, it may represent merely a biotransformation and not ultimate degradation. Indeed partial dechlorination is extremely common. A review of anaerobic dechlorination has been given (Mohn and Tiedje 1992). Under these conditions—and in contrast to aerobic conditions—the more highly substituted compounds are more readily transformed. Partial anaerobic dechlorination of pentachlorophenol has been observed and, since the resulting chlorophenols are degradable under aerobic conditions complete mineralization may be accomplished under alternating anaerobic–aerobic conditions. Pure cultures of Desulfomonile tiedjei are able to carry out dechlorination, strain DCB-1 of polychlorinated phenols (Mohn and Kennedy 1992), and strain DCB-2 of 2,4,6-trichlorophenol and dichlorophenols (Madsen and Licht 1992). Desulfitobacterium hafniense (frappieri) strain PCP-1 that was isolated from a methanogenic consortium can dechlorinate pentachlorophenol to 3-chlorophenol via 2,3,4,5-tetra-, 3,4,5-tri-, and 3,5-dichlorophenol (Bouchard et al. 1996). Reinoculation was, however necessary to achieve >95% removal with high concentrations of pentachlorophenol (Beaudet et al. 1998). In addition, this strain is able to dechlorinate a wide range of polychlorinated aromatic substrates including phenols, catechols, anilines, pentachloronitrobenzene (PCNB), and pentachloropyridine (Dennie et al. 1998).

CONCLUSION There is evidence for the success of bioremediation of chlorophenol-contaminated sites by on-site processes using a number of aerobic microorganisms as inocula, although their activity depended on a number of factors that should be investigated for each site (Miethling and Karlson 1996). However, advantage should be taken of indigenous microflora together with the addition of suitable sorbents. A relatively high degree of mineralization may be attained, although the possibly undesired formation of chloroanisoles could present a potential hazard, together with the fungal-catalyzed polymerization of chlorophenols. The extent to which this is applicable to chlorophenols containing methyl or nitro substituents is apparently unresolved.

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Häggblom MM, RJ Valo (1995) Bioremedation of chlorophenol wastes. In Microbial transformation and Degradation of Toxic Organic Chemicals (Eds LY Young and CE Cerniglia), pp. 389–434. Wiley-Liss, New York. Järvinen KT, ES Melin, JA Puhakka (1994) High-rate bioremediation of chlorophenol-contaminated groundwater at low temperatures. Environ Sci Technol 28: 2387–2392. Laine MM, KS Jörgensen (1996) Straw compost and bioremediated soil as inocula for the bioremediation of chlorophenol contaminated soil. Appl Environ Microbiol 62: 1507–1513. Laine MM, KS Jörgensen (1997) Effective and safe composting of chlorophenol contaminated soil in pilot scale. Environ Sci Technol 30: 371–378. Lamar RT, DM Dietrich (1990) In situ depletion of pentachlorophenol from contaminated soil by Phanerochaete spp. Appl Environ Microbiol 56: 3093–3100. Lamar RT, JW Evans, J Glaser (1993) Solid-phase treatment of a pentachlorophenol-contaminated soil using lignin-degrading fungi. Environ Sci Technol 27: 2566–2571. Madsen T, D Licht (1992) Isolation and characterization of an anaerobic chlorophenol-transforming bacterium. Appl Environ Microbiol 58: 2874–2878. McCarthy DL, AA Claude, SD Copley (1997) In vivo levels of chlorinated hydroquinones in a pentachlorophenoldegrading bacterium. Appl Environ Microbiol 63: 1883–1888. Miethling R, U Karlson (1996) Accelerated mineralization of pentachlorophenol in soil upon inoculation with Mycobacterium chlorophenolicum PCP1 and Sphingomonas chlorophenolica RA2. Appl Environ Microbiol 62: 4361–4366. Mohn WW, JM Tiedje (1992) Microbial reductive dehalogenation. Microbiol Rev 56: 482–507. Mohn WW, KJ Kennedy, (1992) Reductive dehalogenation of chlorophenols by Desulfomonile tiedjei DCB-1. Appl Environ Microbiol 58: 1367–1370. Palm H, J Knuutinen, J Haimi, J Salminen, V Huhta (1991) Methylation products of chorophenols, catechols and hydroquinones in soil and earthworm of sawmill environments. Chemosphere 23: 263–267. Steiert JG, RL Crawford (1986) Catabolism of pentachlorophenol by a Flavobacterium sp. Biochem Biophys Res Commun 141: 1421–1427. Tse KKC, S-L Lo, JWH Wang (2001) Pilot study of in-situ thermal treatment for the remediation of pentachlorophenol-contaminated aquifers. Environ Sci Technol 35: 4910–4915. Valli K, MH Gold (1991) Degradation of 2,4-dichlorophenol by the lignin-degrading fungus Phanerochaete chrysosporium. J Bacteriol 173: 345–352. Valo RJ, MM Häggblom, MS Salkinoja-Salonen (1990) Bioremediation of chlorophenol containing simulated ground water by immobilized bacteria. Water Res 24: 253–258. Xun L, E Topp, CS Orser (1992) Confi rmation of oxidative dehalogenation of pentachlorophenol by a Flavobacterium pentachlorophenol hydroxylase. J Bacteriol 174: 5745–5747.

PART 4 CHEMICAL WASTE The range of compounds in waste or in by-products from the production of chemical is enormous. It includes aliphatic and aromatic and heterocyclic compounds, some of which contain fluorine, chlorine, bromine, nitro, sulfonate, or thiophosphate substituents. A useful structural listing of a wide range of chemicals has been compiled (Hartter 1985; Swoboda-Colberg 1995), and attention is drawn to two less-prominent groups containing the CF3 group, and organic iodine compounds such as the herbicide ioxynil and x-ray contrast agents. Chlorophenols have already been discussed and commercial chlorophenols may contain impurities including chlorinated dioxins and furans. Phenoxyalkanoic acid herbicides are discussed in the section on agrochemicals, and nitro compounds and phosphorothioate esters in later sections. The present discussion concentrates on halogenated aromatic and alicyclic hydrocarbons, although attention is directed also to other structural groups such as benzothiazoles (Reemtsma et al. 1995; Kumata et al. 1996) that have wide industrial application, although less is known of their biodegradability (De Vos et al. 1993). It should be emphasized that accumulation of by-products may present problems so that attention should not be confined to the manufactured product. Illustrative examples include (a) diphenyl sulfone as a by-product from the production of phenol by sulfonation of benzene followed by hydrolysis (Wick and Gschwend 1998), and (b) the diverse compounds containing chlorine and trifluoromethyl substituents resulting from the production of 4-chloro-(trifluoromethyl) (Jaffe and Hites 1985), which is noted later.

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CHLORINATED ALICYCLIC HYDROCARBONS Hexachlorocyclohexane (HCH) isomers were among the organochlorine contaminants at sites contaminated with previous pesticide production (Lang et al. 1992). Although evidence has been presented for the existence of a rich soil microflora at such sites, it appears that many of the substrates may remain even after 30 years (Feidieker et al. 1994). This suggests the existence of one or more limiting factors, of which the degree of bioavailability is probably of major significance, and is consistent with the established difficulty of chemical recovery of HCH from aged contaminated soil (Westcott and Worobey 1985). The situation is additionally complicated by the existence of various isomers of HCH (α,β,γ,δ) each of which behaves differently, and the existence of enantiomers of the α-isomer. The aerobic degradation of HCH isomers has been discussed in Chapter 9, Part 1. The degradation of γ-hexachloro[aaaeee]cyclohexane has been accomplished by several bacteria including Pseudomonas paucimobilis strain UT26 (Nagasawa et al. 1993), and involves a complex sequence of elimination, hydrolytic displacement of chlorine, and ring-fission reactions. It should be noted that the biotransformation to chlorobenzenes as terminal metabolites presents an undesirable alternative. The aerobic degradation of α-hexachloro[aaaaee]cyclohexane by indigenous bacteria in soil slurries has been demonstrated (Bachmann et al. 1988) and, was apparently limited by the rates of soil desorption and intraparticle mass transfer (Rijnaarts et al. 1990). Although the β-isomer (eeeeee) is highly persistent, its susceptibility to transformation under methanogenic conditions in a laboratory column reactor has been studied (Middeldorp et al. 1996). Chlorobenzene and benzene were terminal metabolites, plausibly produced from an intermediate δ-2,3,4,5-tetrachlorocyclohex1-ene. The anaerobic transformation of the structurally related γ-hexachlorocyclohexene and γ-1,3,4,5,6-pentachlorocyclohexene by Clostridium rectum produced 1,2,4-trichlorobenzene and 1,4-dichlorobenzene, respectively (Ohisa et al. 1980). Considerable attention has been directed to the highly toxic polychlorinated bornanes (toxaphene), and derivatives of hexachlorocyclopentadiene (chlordane) that were used as pesticides. Chlordane has been recovered from biota (Buser et al. 1992) and toxaphene components from sediments (Stern et al. 1996), so that these are highly persistent contaminants. For both, the commercial products contain a large number of congeners, and the analysis of chlordane is complicated additionally by the occurrence of enantiomers, some of which have undergone degradation and transformation after discharge. Toxaphene is a complex mixture of compounds prepared by chlorinating camphene, and contains several hundred polychlorinated bornanes. Although experiments with Sulfurospirillum (Dehalospirillum) multivorans have demonstrated partial dechlorination (Ruppe et al. 2003, 2004), only limited attention appears to have been directed to waste from locations at which these products were manufactured.

CHLORINATED AND BROMINATED AROMATIC HYDROCARBONS The structural range of industrially important representatives of these groups is enormous, and includes chlorobenzenes (solvents), polychlorinated biphenyls (PCBs) (hydraulic and insulating fluids), and polybrominated biphenyls and diphenyl ethers (flame retardants). There is widespread concern over both the persistence and the potential toxicity of all these compounds, and sites that have become contaminated during their production represent a threat both to the environment and to human health. Pathways for the aerobic bacterial degradation of chlorobenzenes and chlorobiphenyls, and their brominated analogs have been discussed in Chapter 9, Part 1.

CHLOROBENZENES Although the possibly incompatible pathways for the degradation of chlorobenzene and toluene has been noted, a strain of Ralstonia sp. from a site contaminated with chlorobenzene contained the genes for both chlorocatechol degradation and the dioxygenase system for the degradation of benzene/toluene (van der Meer et al. 1998). The evolution of this strain resulted in a natural lowering

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in the groundwater concentration. Aerobic organisms capable of degrading chlorobenzenes with less than five chlorine atoms have been described and the pathways elucidated: the metabolism can be initiated by dioxygenation without elimination of a chlorine atom from 1,2,4-trichlorobenzene or with elimination of a single chlorine atom from 1,2,4,5-tetrachlorobenzene (Sander et al. 1991). The possible application of the soluble methane monooxygenase of Methylosinus trichosporium to the bioremediation of 1,2,3-trichlorobenzene is attended by three negative factors: (a) the formation of the toxic 2,3,4- and 3,4,5-trichlorophenol, (b) the requirement of a substrate such as formate, and (c) the permanent inactivation of the cells by the substrate (Sullivan and Chase 1996). The anaerobic dechlorination of hexachlorobenzene has been described in anaerobic mixed cultures supplemented with electron donors including lactate, ethanol, or glucose (Holliger et al. 1992): successive and partial dechlorination produced 1,2,4- and 1,3,5-trichlorobenzenes, while the 1,2,3-trichlorobenzene was further dechlorinated. The partial dechlorination of 1,2,3,4-tetra-, 1,2,3,5-tetra-, and pentachlorobenzene has been examined in a methanogenic mixed culture using lactate as electron donor (Middeldorp et al. 1997), and sterile Rhine River sand was needed to maintain dechlorination activity for unresolved reasons.

DDT There has been renewed interest in DDT in spite of the fact that its use has been banned in many countries for several years. The degradation of DDT has been discussed in Chapter 7, Part 3, and attention has been directed to the apparently recalcitrant DDE (Quensen et al. 1998) and DDA (Heberer and Dünnbier 1999). Uptake and metabolism of p,p′-DDT and the isomeric o,p′-DDT have been examined in a range of plants, and the results illustrate a number of important issues. 1. One study used weathered contaminated soil that had been exposed to DDT for 3 years and had lain undisturbed since 1950 (Lunney et al. 2004). Three groups of plants were used, zucchini and pumpkin (Cucurbita pepo hybrids), alfalfa (Medicago sativa), and the grasses, tall fescue (Festuca arundinacea) and rye grass (Lolium multiflorum), at two levels of exposure. Accumulation was greatest in the roots and shoots for both strains of Cucurbita pepo, whereas for the others accumulation was essentially restricted to the roots. The grasses are distinguished by the development of extensive root systems, whereas the Curcubita pepo have an extensive above-ground biomass and leaf surfaces that could have a positive effect in translocating the contaminant from the roots to the shoots. It was suggested that in addition, plant exudates might contribute to uptake. 2. The metabolism of DDT in aqueous solution was examined in the aquatic plant Elodea canadensis, and in stems and leaves of the perennial vine Pueraria thunbergiana (Garrison et al. 2000). Use of [14C]-labeled DDT confirmed the uptake of DDT that was strongly bound to the plant material. It was shown that DDT was transformed to DDD by reductive dechlorination, and significant aspects emerged: (a) the reaction took place with irradiated samples that contained no viable microorganisms, which were not therefore directly involved, and (b) dead plants of Elodea maintained reductive capability. It was concluded that the observed reduction was therefore abiotic, although it was dependent on some unknown factor that was present in the plants. This pathway of transformation to DDD in Elodea canadensis was confirmed in a study using the additional aquatic plants, duckweed (Spirodela oligorrhiza) and parrot feather (Mariophyllum aquaticum) (Gao et al. 2000).

POLYCHLORINATED BIPHENYLS There is an enormous literature on the degradation of PCBs, and it is not intended to attempt to cover this. The degradation of PCBs is exacerbated not only by the number of congeners, but by the fact that heavily ortho-substituted congeners occur in enantiomeric forms as a result of restricted

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rotation between the rings. Attention is directed to reviews on anaerobic dechlorination (Bedard and Quensen 1995), and to others that include a summary of aerobic degradation (Bedard 1990; Sylvestre 1995). The last also includes valuable comments on the regulation of PCB degradation and the significance of chlorobenzoate degradation. As a background to a discussion of bioremediation technologies, a few simplistic remarks are given here on the metabolic pathways used for the degradation of PCBs that has been discussed in detail in Chapter 9, Part 1. It should be emphasized that there are very considerable differences among the congeners in their susceptibility to degradation, in details of the mechanisms whereby this is accomplished, and among the various organisms. No attempt is, however, made to discuss here the details of these important differences. One important parameter is temperature, that is of particular significance for naturally occurring mixed cultures of organisms in the natural environment, since temperature may result in important changes in the composition of the microbial flora as well as on the rates. Some illustrative examples of the importance of temperature under both aerobic and anaerobic situations include the following: 1. An anaerobic sediment sample was incubated with 2,3,4,6-tetrachlorobiphenyl at various temperatures between 4 and 66°C (Wu et al. 1997). The main products were 2,4,6- and 2,3,6-trichlorobiphenyl and 2,6-dichlorobiphenyl: the first of these was produced maximally and discontinuously at 12 and 34°C, the second maximally at 18°C, and the third was dominant from 25 to 30°C. Dechlorination was not observed above 37°C. 2. Sediment samples from a contaminated site were spiked with Arochlor 1242 and incubated at 4°C for several months (Williams and May 1997). Degradation by aerobic organisms in the upper layers of the sediment occurred with the selective production of di- and trichlorobiphenyls—although not in those at more than 15 mm below the surface. Some congeners were not found including the 2,6- and 4,4′-dichlorobiphenyls and a wider range of trichlorobiphenyls which were presumably further degraded. A number of aerobic organisms are able to degrade various congeners of PCB and has been discussed in Chapter 9, Part 1. Most have been isolated by enrichment with biphenyl as sole carbon and energy source. These include both Gram-negative strains belonging to the genera Alcaligenes and Pseudomonas, and Gram-positive strains of the genus Rhodococcus. Growth and the ability to carry out dechlorination may be accomplished by bacteria growing not only at the expense of biphenyl that was used for enrichment of the organisms, but also at the expense of typical plant metabolites that are noted in Chapter 12. In an ingenious application, the recovery of chlorobenzoates has been used to demonstrate aerobic degradation under natural conditions (Flanagan and May 1993). Protoanemonin can be formed from 4-chlorocatechol that is an intermediate in the degradation if 4-chlorobiphenyl (Blasco et al. 1995). This has an adverse effect on PCB-degrading bacteria in soil (Blasco et al. 1997), and may therefore present a significant problem in the bioremediation of PCB-contaminated sites. Conversely, naturally occurring compounds may be able to induce the enzymes for PCB metabolism. For example, effective degradation by cometabolism of a number of congeners in Arochlor 1242 was induced in Arthrobacter sp. strain B1B by the monoterpenoid carvone that was not used as a growth substrate, although it was toxic at high concentrations (>500 mg/L) (Gilbert and Crowley 1997). Other structurally related compounds of terpenes including limonene, p-cymene, and isoprene were also effective. Aerobic Conditions A valuable overview (Adriaens and Vogel 1995) includes a range of chlorinated aromatic compounds and provides illustrative engineering designs. In the light of the microbiological background that has already been presented, a number of different strategies can be distinguished. All of these have been applied to, and evaluated in contaminated soils or sediments.

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1. Contaminated soil has been inoculated with bacteria of established capacity for degradation of chlorobenzoates (Hickey et al. 1993). In the presence of added biphenyl, mineralization of PCBs was shown, although it was pointed out that there may exist incompatibility between the production of chlorocatechols from chlorobenzoates and their inhibition of dihydroxybiphenyl-2,3-dioxygenase that catalyzes the ring fission of many PCBs. 2. Experiments using Pseudomonas testosteroni strain B-356 in microcosm systems revealed the necessity of adding biphenyl to promote degradation of the tetrachlorinated congeners of Arochlor 1242, and the effectiveness of repeated inoculation (Barriault and Sylvestre 1993). 3. In situ stimulation of aerobic PCB degradation has been shown in Hudson River sediments (Harkness et al. 1993). Biodegradative capacity of the indigenous organisms was enhanced by the addition of biphenyl and inorganic nutrients, although repeated inoculation with an organism having established capability to degrade PCBs did not apparently improve the degree of biodegradation. 4. As for PAHs, attempts have been made to increase bioavailability by use of surfactants, and a complex picture has again developed (Fava and Di Gioia 1998). Triton-100 exerted both positive and negative effects: in soil slurries even though it was not metabolized by the soil microflora, it adversely affected the degradation of chlorobenzoate intermediates, whereas in fixed-bed reactors, depletion of PCBs was enhanced. 5. It is well established that the enzymes for the aerobic degradation of PCBs can be induced by biphenyl. Efforts have therefore been made to find naturally occurring metabolites that would function as inducers, and whether plants are able to exert a beneficial effect. The putative role of terpenoids in inducing the degradation of PCBs has been noted in Chapter 9, Part 2 and Chapter 12. 6. An important study examined soil in the neighborhood of trees at a site historically contaminated with Arochlor 1260 (Leigh et al. 2006). Organisms able to degrade PCBs were isolated by sequential screening using biphenyl and 4-bromobiphenyl, followed by analysis for their ability to degrade PCBs. Highest numbers of PCB-degrading bacteria were found near mature stands of Pinus nigra (Austrian pine) and Salix caprea (willow). Strains with degradative activity comparable to that of Burkholderia xenovorans LB400 and Rhodococcus sp. strain RHA1 were found in soil from the willow. The most active strains displayed high similarities to Rhodococcus ruber, Rh. erythreus, and Pseudomonas frederiksbergensis, but activity was also found in strains of Williamsia and Luteibacter that have not hitherto been reported to degrade PCBs. These results clearly underscore the importance of plant metabolites in supporting the growth of bacteria with important degradative potential. Anaerobic Conditions Anaerobic dechlorination has been demonstrated under a number of conditions, and a complex pattern of dechlorination among the congeners has been observed (Bedard and Quensen 1995). There are also varying, and sometimes conflicting, views on the role of sulfate. 1. The anaerobic dechlorination of PCBs has been extensively studied both in microcosms and in field samples from heavily contaminated sites in the United States. Three main patterns have been found—N that removed flanked meta-chlorines, P that removed para-chlorines, and LP that removed unflanked para-chlorines (Bedard et al. 1998). By contrast, ortho-chlorines were more recalcitrant. These experiments, which have been discussed in Chapter 9, Part 2, laid the foundation for analysis of a field situation and an appreciation of the effect of long-term exposure of contaminated lake sediment (Magar et al. 2005a,b). Substantial dechlorination took place in buried sediment cores (35–40 cm) compared with the surface sediment cores (0–5 cm). Although there were some variations among the cores, the di- and trichlorinated biphenyls were produced at the expense of the

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more highly chlorinated congeners. There was preferential removal of meta- and parachlorine substituents, whereas the ortho-substituted congeners were essentially recalcitrant even after burial for 50 years, which was consistent with previous observations in laboratory experiments. Important conclusions were that (a) rates of dechlorination were substantially less than those expected on the basis of some laboratory experiments, and (b) sediments exposed near the surface were dechlorinated only to a very limited extent. 2. Anaerobic strategies are complicated by the different susceptibility of the various congeners to dechlorination and by the effect of electron acceptors especially sulfate. Ingenious alternatives have used addition of specific PCB congeners that are more readily dechlorinated to “prime” dechlorination at specific positions (Bedard and Quensen 1995), and the use of dibromo biphenyls in the presence of malate to stimulate dechlorination of the hexachloroto nanochlorobiphenyls (Bedard and Quensen 1995). The use of brominated biphenyls to induce dechlorination of highly chlorinated biphenyls has been systematically examined. Di- and tribromobiphenyls were the most effective in priming dechlorination including the hepta-, hexa-, and pentachloro congeners, and they were subsequently reduced to biphenyl (Bedard et al. 1998). 2,6-Dibromobiphenyl stimulated the growth of anaerobes that effectively dechlorinated hexa-, hepta-, octa-, and nanochlorobiphenyls over the temperature range from 8 to 30°C (Wu et al. 1999). Other compounds have been examined as “primers” for dechlorination of hexa- to nonachloro congeners (Deweerd and Bedard 1999). For example, 4-bromobenzoate was effective—though less so that 2,6-dibromobiphenyl—whereas the chlorobenzoates that are metabolites of aerobic degradation of PCBs were ineffective. The positive effect of brominated biphenyls in “priming” the anaerobic dechlorination of PCBs has also been encountered in the dechlorination of octachlorodibenzo[1,4]dioxin to the 2,3,7,8 congener that was induced by 2-bromodibenzo[1,4]dioxin (Albrecht et. al 1999).

POLYBROMINATED BIPHENYLS AND DIPHENYL ETHERS Highly brominated biphenyls and diphenyl ethers have been manufactured as flame retardants, and the diphenyl ethers have apparently become widespread environmental contaminants (references in Sellström et al. 1998). One reason for concern is the formation of polybrominated dibenzofurans and dibenzo[1,4]dioxins (though not 2,3,7,8-tetrabromodibenzo[1,4]dioxin) during subsequent incineration (Buser 1986). Mixed cultures of organisms isolated from sediments contaminated with PCBs and PBBs were shown to debrominate PBBs under anaerobic conditions (Morris et al. 1992), and the dominant congener—2,4,5,2′,4′,5′-hexabromobiphenyl—could be successively debrominated to 2,2′-dibromobiphenyl. In contrast, in sediments from the most heavily contaminated site containing contaminants in addition to PBBs, very little debromination occurred, and the recalcitrance was attributed to the toxicity of the other contaminants (Morris et al. 1993). Debromination of decabromodiphenyl ether has been described and, as for PCB, the rate was increased by the addition of primers (Gerecke et al. 2005). Octa- and nanobrominated congeners were produced, predominantly by loss of bromine at the meta and para positions, and not appreciably at the ortho position. An alternative procedure used two anaerobes that were able to dechlorinate chlorinated ethenes, and the results showed significant differences between them. One of them debrominated decabromodiphenyl ether to octa and hepta congeners, whereas the other that was unable to debrominate decabromodiphenyl ether, carried out debromination of the octa congeners to the hexa (154) and subsequently to the penta (99) and tetra (49) + tetra (47) congeners (He et al. 2006).

CHLORINATED DIBENZODIOXINS These are probably ubiquitous in the environment, and their biodegradability has been extensively examined. This will not be discussed in detail here, and attention will be drawn merely to a few investigations that illustrate important aspects.

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Chlorinated dioxins have been recovered from sediments contaminated with industrial discharge and from atmospheric deposition (Götz et al. 1993; Macdonald et al. 1992; Evers et al. 1993). Attention has been directed to anaerobic dechlorination processes (Beurskens et al. 1995; Adriaens et al. 1995). It has been suggested that microbial dechlorination of cell-partitioned 2,3,7,8tetrachlorodibenzo[1,4]dioxin in aged sediments was as effective as that in freshly spiked sediment (Barkovskii and Adriaens 1996). This suggests that the kinetics of transport into the cells is a cardinal determinant, although the extent to which hydrophobic xenobiotics in sediments are partitioned into the indigenous microflora has not been generally established. Chlorinated dioxins occur in atmospheric deposition (Koester and Hites 1992), and will thereby enter the terrestrial environment and watercourses. The degradation of tetrachloro- through octachlorodibenzo[1,4]dioxins has been examined in low-nitrogen medium by Phanerochaete sordida YK-624 (Takada et al. 1996). All the compounds were extensively degraded, and the ring fission of 2,3,7,8-tetra- and octachlorodibenzo[1,4]dioxin produced 4,5-di- and tetrachlorocatechol. These results established important evidence for the biodegradability of even highly chlorinated dibenzodioxins.

ORGANOFLUORO COMPOUNDS These have found a wide range of application including use as, for example, agrochemicals (Cartwright 1994), liquid crystals (Inoi 1994), fluorine-containing dyes (Herd 1994; Engel 1994), plastics (Feiring 1994), elastomers (Logothetis 1994), and chemotherapeutic agents (Edwards 1994). The plethora of structures is a tribute to the ingenuity of the synthetic chemist, and both aliphatic and aromatic structures are widely represented. Examples include diphenyl ether herbicides containing the CF3 group (Lee et al. 1995), and fluoroaromatic compounds used as pharmaceuticals such as the fluoroquinolone carboxylic acids that are used in substantial amounts in large-scale animal husbandry (Nowara et al. 1997). A few examples of the existence of fluorinated organic compounds are used as illustration of a problem area. 1. Sediments containing waste from the production of 4-chloro-(trifluoromethyl)benzene contained a number of compounds containing both chlorine and trifluoromethyl substituents. These included a benzophenone, a difluorodiphenylmethane, and several biphenyls (Jaffe and Hites 1985), although the more widespread occurrence of these compounds is unknown. This is important since, whereas pathways for the degradation of fluorophenols and fluorobenzoates are well established, the CF3 group appears to be resistant to both chemical and microbiological degradation. Microbial reactions brought about the degradation of 3- and 4-trifluoromethylbenzoate to 7,7,7-trifluoro-hepta-2,4-diene-6-one carboxylate that could, however, be degraded photochemically with the loss of fluoride (Taylor et al. 1993). 2. A series of dibenzo[1,4]dioxins bearing nitro and trifluoromethyl substituents in one ring and chlorine in the other was synthesized for comparison with impurities in commercial samples of the lampricide 3-trifluoromethyl-4-nitrophenol, and to assess their ability to induce MFO activity (Hewitt et al. 1998). Although they appeared to be relatively weak inducers, there were other unidentified components that were highly active. 3. Polyfluorinated carboxylic acids and sulfonic acids have been used extensively, for example, in fire-retardant foams, as surfactants under extreme conditions, and as additives to fabrics and domestic products. What is more disturbing is their ubiquity in a wide range of environmental samples including humans. There is no evidence for their biodegradation, although photochemical procedures have been successful in bringing about their partial degradation (Hori et al. 2004a,b; 2005a,b).

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CONCLUSION There are a number of factors that contribute to the difficulty in the bioremediation of waste from chemical production. These include the enormous range of chemical structures and this is exacerbated by the frequent occurrence of structurally related compounds that have different susceptibilities to degradation. A major problem is exemplified by the complex mixture in commercial preparations of PCBs that includes not only a range of congeners but also polychlorinated terphenyls (Wester et al. 1996). In addition, most sites have been subjected to historical contamination so that aging that results in reduced bioavailability may present a serious problem. Although several strategies have been used, most often with bacteria, increasing interest has been directed to the use of plants. This is illustrated by PCBs where there is unequivocal evidence for successful bioremediation of PCBs. Current research is directed to understanding the basis of the biodegradability of specific congeners, and to the possibility of developing strategies to overcome this. All these considerations apply to contaminants including PBBs, PCDDs, and aromatic compounds containing CF3 groups for which cardinal aspects of the basic microbiology remain to be developed.

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Hori H, A Yamamoto, E Hayakawa, S Taniyasu, S Kutsuna, H Kiatagawa, R Arakawa (2005a) Efficient decomposition of environmentally persistent perfluorocarboxylic acids by use of persulfate as a photochemical oxidant. Environ Sci Technol 39: 2383–2388. Hori H, A Yamamoto, S Kutsuna (2005b) Efficient photochemical decomposition of long-chain perfluorocarboxylic acids by means of an aqueous/liquid CO2 biphasic system. Environ Sci Technol 39: 7692–7697. Hori H, E Hayakawa, H Einaga, S Kutsuna, K Koike, T Ibusuki, H Koatagawa, R Arakawa (2004a) Decomposition of environmentally persistent perfluorooctanoic acid in water by photochemical approaches. Environ Sci Technol 38: 6118–6124. Hori H, E Hayakawa, K Koike, H Einaga, T Ibusuki (2004b) Decomposition of nonafluoropentanoic acid by heteropolyacid photocatalyst H3PW12O40 in aqueous solution. J Mol Cat A 211: 35–41. Inoi T (1994) Fluorinated liquid crystals. In Organofluorine Chemistry Principles and Commercial Applications (Eds RE Banks, BE Smart, JC Tatlow), pp. 263–286. Plenum Press, New York. Jaffe R, RA Hites (1985) Identification of new, fluorinated biphenyls in the Niagara River Lake Ontario area. Environ Sci Technol 19: 736–740. Koester CJ, RA Hites (1992) Wet and dry deposition of chlorinated dioxins and furans. Environ Sci Technol 26: 1375–1382. Kumata H, H Takada, N Ogura (1996) Determination of 2-4-morpholinyl benzothiazole in environmental samples by a gas chromatograph equipped with a flame photometric detector. Anal Chem 68: 1976–1981. Lang E, H Viedt, J Egestorff, HH Hanert (1992) Reaction of the soil microflora after contamination with chlorinated aromatic compounds and HCH. FEMS Microbiol Ecol 86: 275–282. Lee HJ, MV Duke, JH Nirk, M Yamamoto, SO Duke (1995) Biochemical and physiological effects of benzheterocycles and related compounds. J Agric Food Chem 43: 2722–2727. Leigh MB, P Prouzová, M Macková, T Macek, DP Nagle, JS Fletcher (2006) Polychlorinated biphenyl (PCB)-degrading bacteria associated with trees in a PCB-contaminated site. Appl Environ Microbiol 72: 2331–2342. Logothetis AL (1994) Fluoroelastomers. In Organofluorine Chemistry Principles and Commercial Applications (Eds RE Banks, BE Smart, JC Tatlow), pp. 373–396. Plenum Press, London. Lunney AA, BA Zeeb, KJ Reimer (2004) Uptake of weathered DDT in vascular plants: potential for phytoremediation. Environ Sci Technol 38: 6147–6154. Macdonald RW, WJ Cretney, N Crewe, D Paton (1992) A history of octachlorodibenzo-p-dioxin, 2,3,7,8tetrachlorodibenzofuran, and 3,3′,4,4′-tetrachlorobiphenyl contamination in Howe Sound, British Columbia. Environ Sci Technol 26: 1544–1550. Magar VS, GW Johnson, RFC Brenner, JF Quensen, EA Foote, G Durell, JA Ickes, C Peven-Mccarthy (2005a) Long-term recovery of PCB-contaminated sediments at the Lake Hartwell superfund site: PCB dechlorination. 1. End-member characterization. Environ Sci Technol 39: 3538–3547. Magar VS, RFC Brenner, GW Johnson, JF Quensen (2005b) Long-term recovery of PCB-contaminated sediments at the Lake Hartwell superfund site: PCB dechlorination. 2. Rates and extent. Environ Sci Technol 39: 3548–3554. Middeldorp PJ, M Jaspers, AJB Zehnder, G Schraa (1996) Biotransformation of α−, β−, γ−, and δ-hexachlorocyclohexane under methanogenic conditions. Environ Sci Technol 30: 2345–2349. Middeldorp PJM, J de Wolf, AJB Zehnder, G Schraa (1997) Enrichment and properties of a 1,2,4-trichlorobenzene-dechlorinating microbial consortium. Appl Environ Microbiol 63: 1225–1229. Morris PJ, JF Quensen, JM Tiedje, SA Boyd (1992) Reductive debromination of the commercial polybrominated biphenyl mixture Firemaster BP6 by anaerobic microorganisms from sediments. Appl Environ Microbiol 58: 3249–3256. Morris PJ, JF Quensen, JM Tiedje, SA Boyd (1993) An assessment of the reductive debromination of polybrominated biphenyls in the Pine River reservoir. Environ Sci Technol 27: 1580–1586. Nagasawa S, R Kikuchi, Y Nagata, M Takagi, M Matsuo (1993) Aerobic mineralization of γ-HCH by Pseudomonas paucimobilis UT26. Chemosphere 26: 1719–1728. Nowara A, J Burhenne, M Spiteller (1997) Binding of fluoroquinolone carboxylic acid derivatives to clay minerals. J Agric Food Chem 45: 1459–1463. Ohisa N, M Yamaguchi, N Kurihara (1980) Lindane degradation by cell-free extracts of Clostridium rectum. Arch Microbiol 125: 221–225. Quensen JF, SA Mueller, MK Jain, JM Tiedje (1998) Reductive dechlorination of DDE to DDMU in marine sediment microcosms. Science 280: 722–724. Reemtsma T, O Fiehn, G Kalnowski, M Jekel (1995) Microbial transformations and biological effects of fungicide-derived benzothiazoles determined in industrial wastewater. Environ Sci Technol 29: 478–485.

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Rijnaarts HHM, A Bachmann, JC Jumelet, AJB Zehnder (1990) Effect of desorption and intraparticle mass transfer on the aerobic biomineralization of alpha-hexachlorocyclohexane in a contaminated calcareous soil. Environ Sci Technol 24: 1349–1354. Ruppe S, A Neumann, E Braekevalt, GT Tomy, GA Stern, KA Maruya, W Vetter (2004) Anaerobic transformation of compounds of technical toxaphene. 2. Fate of compounds lacking geminal chlorine aroms. Environ Toxicol Chem 23: 591–598. Ruppe S, A Neumann, W Vetter (2003) Anaerobic transformation of compounds of technical toxaphene. I. Regiospecific reaction of chlorobornanes with geminal chlorine aroms. Environ Toxicol Chem 22: 2614–2621. Sander P, R-M Wittich, P Fortnagel, H Wilkes, W Francke (1991) Degradation of 1,2,4-trichloro- and 1,2,4,5tetrachlorobenzene by Pseudomonas strains Appl Environ Microbiol 57: 1430–1440. Sellström U, A Kierkegaard, C de Witt, B Jansson (1998) Polybrominated diphenyl ethers and hexabromocyclododecane in sediment and fish from a Swedish river. Environ Toxicol Chem 17: 1065–1072. Stern GA, MD Loewen, BM Miskimmin, DCG Muir, JB Westmore (1996) Characterization of two major toxaphene components in treated lake sediment. Environ Sci Technol 30: 2251–2258. Sullivan JP, HA Chase (1996) 1,2,3-Trichlorobenzene transformation by Methylosinus trichosporium OB3b expressing soluble methane monooxygenase. Appl Microbiol Biotechnol 45: 427–433. Swoboda-Colberg NG (1995) Chemical contamination of the environment: sources, types, and fate of synthetic organic chemicals. In Microbial transformation and degradation of toxic organic chemicals (Eds LYYoung and CE Cerniglia), pp. 27–74. Wiley-Liss, New York. Sylvestre M (1995) Biphenyl/chlorobiphenyls catabolic pathway of Comamonas testosteroni B-356: prospect for use in bioremediation. Int Biodet Biodeg 35: 189–211. Takada S, M Nakamura, T Matsueda, R Kondo, K Sakai (1996) Degradation of polychlorinated dibenzo-pdioxins and polychlorinated dibenzofurans by the white rot fungus Phanerochaete sordida YK-624. Appl Environ Microbiol 62: 4323–4328. Taylor BF, JA Amador, HS Levinson (1993) Degradation of meta-trifluoromethylbenzoate by sequential microbial and photochemical treatments. FEMS Microbiol Lett 110: 213–216. Van der Meer JR, C Werlen, SF Nishino, JC Spain (1998) Evolution of a pathway for chlorobenzene metabolism leads to natural attenuation in contaminated groundwater. Appl Environ Microbiol 64: 4185–4193. Westcott ND, BL Worobey (1985) Novel solvent extraction of lindane from soil. J Agric Food Chem 33: 58–60. Wester PG, J de Boer, UAT Brinkman (1996) Determination of polychlorinated terphenyls in aquatic biota and sediment with gas chromatography/mass spectrometry using negative chemical ionization. Environ Sci Technol 30: 473–480. Wick LY, PM Gschwend (1998) Source and chemodynamic behaviour of diphenyl sulfone and ortho- and para-hydroxybiphenyl in a small lake receiving discharges from an adjacent superfund site. Environ Sci Technol 32: 1319–1328. Williams WA, RJ May (1997) Low-temperature microbial aerobic degradation of polychlorinated biphenyls in sediment. Environ Sci Technol 31: 3491–3496. Wu Q, DL Bedard, J Wiegel (1997) Effect of incubation temperature on the route of microbial reductive dechlorination of 2,3,4,6-tetrachlorobiphenyl in polychlorinated biphenyl (PCB)-contaminated and PCB-free freshwater sediments. Appl Environ Microbiol 63: 2836–2843. Wu Q, DL Bedard, J Wiegel (1999) 2,6-Dibromobiphenyl primes extensive dechlorination of Arochlor 1260 in contaminated sediment at 8–30°C by stimulating growth of PCB-dehalogenating microorganisms. Environ Sci Technol 33: 595–602.

PART 5 AGROCHEMICALS As for waste from the production of chemicals, the array of structures represented by agrochemicals is truly enormous. Only some illustrative examples are provided, and it is important to emphasize that not only the original compound, but also potential metabolites should be considered. The pathways for biodegradation of many of the structures have been presented in Chapter 9 and reference should be made to these for details. There is increased interest in the degradation of agrochemicals after application, and abiotic reactions including photochemical degradation that are important on the soil surface are discussed in Chapter 1.

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PHENOXYALKANOIC ACIDS The pathway for their degradation that involves the intermediate formation of chlorophenols is well established, and several groups of organisms able to do so have been isolated. Kinetic parameters for the rate of degradation and for sorption of 2,4-dichlorophenoxyacetic acid (2,4-D) (Estrella et al. 1993) have been carried out, and showed higher values for the maximum growth rate constants in column experiments than in batch experiments. Experiments on the bioremediation of 2,4-Dcontaminated soil using Pseudomonas cepacia strain BRI 6001 were promising, and indicated the critical role of the cell density of the inoculum added to nonsterile soil (Comeau et al. 1993). However, the bacterial degradation of 2,4-D sorbed to sterile soil has been shown to be extremely slow (Ogram et al. 1985), and the rates in soil with a low organic content were greater than in an organic-rich soil (Greer and Shelton 1992). The biodegradation of phenoxypropion-2-oic acids is less widely established, and mecoprop [2-(2-methyl-4-chlorophenoxy)propionic acid] has consistently been found in municipal landfill leachates (Gintautas et al. 1992). Whereas in pure cultures of bacteria, only one enantiomer of 2-(2-methyl-4-chlorophenoxy)propionic acid—the (S)-(–)enantiomer—may be degraded (Tett et al. 1994), in agricultural soil treated with a herbicide mixture containing 2-(2,4-dichlorophenoxy)propionic acid, both enantiomers were degraded with the opposite preference (Garrison et al. 1996). The same preference for degradation of racemic 2-(2-methyl4-chlorophenoxy)propionic acid was observed in Sphingomonas herbicidovorans by Zipper et al. (1996) who made the important observation that cells grown with either of the pure enantiomers preferentially degraded the corresponding enantiomer. Current formulations contain, however, only the biologically active and more readily biodegradable enantiomer. Additional comments on the enantiomerization of 2-phenoxypropionic acids (Buser and Müller 1997) have been given.

CHLORINATED ANILINES These are the starting compounds for production of a range of agrochemicals (insecticides, herbicides, and fungicides) including anilides, carbamates, and ureas. In addition, the chloroanilines produced by their hydrolysis are widely distributed, and are notorious for their association with and incorporation into humic substances (Bollag and Loll 1983). The mechanism of this interaction has been examined in model enzyme-mediated reactions with ferulic acid (Tatsumi et al. 1994), and it has been shown that oligomerization of 4-chloroaniline mediated by oxidoreductases may produce 4,4′-dichloroazobenzene and 4-chloro-4′-aminodiphenyl as well as trimers and tetramers (Simmons et al. 1987). These associations are also expected for the chlorinated anilines produced from the fungicide PCNB. From a purely microbiological view, there are some important additional issues. 1. Although aerobic bacteria that degrade monochloroanilines have been isolated, the biodegradation of dichloroanilines appears to be more exceptional (Brunsbach and Reineke 1993). In particular, the regulation of chloroaniline degradation should be taken into account, including the possible requirement for additional carbon sources. The anaerobic dechlorination of chloroanilines has been discussed in Chapter 9, Part 2, and the pathway for the anaerobic degradation of aniline itself is carried out by carboxylation followed by deamination. 2. Biotransformation to acetanilides has been demonstrated, although a strain of Pseudomonas putida stimulated by aniline was able to degrade 3,4-dichloroaniline (You and Bartha 1982). The formation of 3,3′,4,4′-tetrachloroazobenzene, 1,3-bis(3,4-dichlorophenyl)triazine, and 3,3′,4,4′-tetrachlorobiphenyl from 3,4-dichloroaniline by Escherichia coli in the presence of nitrate plausibly involves intermediate formation of the diazonium compound by reaction of the amine with nitrite (Corke et al. 1979). The chloroaniline component of lignin complexes is slowly mineralized by Phanerochaete chrysosporium, though as effectively as the free compound (Arjmand and Sandermann 1985). The mineralization of 3,4-dichloroaniline by P. chrysosporium proceeds sequentially by conjugation with 2-ketoglutaryl CoA to produce the amide and N-(3,4-dichloro)succinamide before

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mineralization (Sandermann et al. 1998). Negative aspects of fungal metabolism include (a) the transformation of 3,4-dichloroaniline by white-rot fungi (Pieper et al. 1992), which may produce the toxic 3,3′,4,4′-tetrachloroazobenzene that is absorbed from plant roots (Lee and Kyung 1995) and (b) facilitated association with humic and fulvic acids. 3. Chlorinated anilines are produced by the reduction of PCNB that is used as a fungicide against a variety of commercial crops. The transformation of PCNB has been examined with a methanogenic enrichment culture from contaminated sediment, although this contained neither PCNB nor its reduction product pentachloroaniline (PCNA). The culture not only reduced the initial PCNB, but also carried out sequential dechlorination to tetra-, tri-, dichloroanilines, and ultimately to 3- and 4-chloroaniline (Tas and Pavlostathis 2005). 4. The removal of 2,4-, 2,6-dichloroaniline, and 2,6-dichloro-4-methylaniline at an aged contaminated site was accomplished by the fungi Allescheriella sp., Stachybotrys, and Phlebia, although their ultimate fate was not determined (D’Annibale et al. 2006). PCNA was removed only by the second of these. 5. A strain that is able to degrade 2-chloro-N-isopropylacetanilide was successful in removing this from soil (Martin et al. 1995). It is notable that this strain was unable to grow with either aniline or phenol. The prognosis for bioremediation of residues containing polychlorinated anilines and their impurities is rather discouraging. Important factors include (a) limited biodegradability, (b) restricted bioavailability, and (c) the possible formation of oligomeric coupling products.

TRIAZINES 1,3,5-Triazines A number of substituted triazines are used as herbicides, and their biodegradation has been discussed in Chapter 10, Part 1. Treatment of soil contaminated with atrazine (2-chloro-4-(ethylamino)-6isopropylamino-1,3,5-triazine) illustrated a number of significant features. Although the soil that was used had the potential for degradation, a laboratory experiment with Pseudomonas sp. strain ADP that had an established potential for atrazine degradation revealed important limitations. There was a substantial decline in the numbers of Pseudomonas sp. strain ADP and only limited mineralization. Supplementation with citrate or succinate increased the survival of the strain, and successful mineralization was dependent on the preservation of a carbon/nitrogen ratio >10 (Silva et al. 2004). The last would apply generally to substrates with a low C/N ratio such as triazines. 1,2,4-Triazines The pathway for the degradation of metribuzin that has been used as a herbicide has been established, and attention has been directed to the occurrence of the metabolites in leachate from agricultural soil in Denmark. Although the parent metribuzin was not found, the metabolites produced by loss of the N-amino group and the thiomethyl groups were produced and contributed to groundwater contamination (Kjaer et al. 2005). 2,4-Dintro-6-sec-Butylphenol (Dinoseb) This was previously used as a herbicide, and attention has been directed to its degradation in storage areas or where it has been spilled. A strain of Clostridium bifermentans KMR-1 (that is protected by a U.S. patent) was unable to use dinoseb as carbon or energy source. In the presence of a starch extract, however, a low level of transformation was observed, and the products could subsequently be mineralized by aerobic bacteria (Hammill and Crawford 1996). These observations have been extended to the remediation of soil slurries from a contaminated site by adding phosphate and starch waste that achieved anaerobic conditions, and inoculation with a culture from a pilot-scale

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experiment. This treatment avoided the formation of undesirable polymeric compounds produced by partial reduction (Kaake et al. 1992; Roberts et al. 1993).

REFERENCES Arjmand M, H Sandermann (1985) Mineralization of chloroaniline/lignin conjugates and of free chloroanilines by the white-rot fungus Phanerochaete chrysosporium. J Agric Food Chem 33: 1055–1060. Bollag J-M, MJ Loll (1983) Incorporation of xenobiotics into soil humus. Experientia 39: 1221–1231. Brunsbach FR, W Reineke (1993) Degradation of chloroanilines in soil slurry by specialized organisms. Appl Microbiol Biotechnol 40: 402–407. Buser H-R, MD Müller (1997) Conversion reactions of various phenoxyalkanoic acid herbicides in soil. 2. Elucidation of the enantiomerization process of chiral phenoxy acids from incubation in a D2O/soil system. Environ Sci Technol 31: 1960–1967. Comeau Y, CW Greer, R Samson (1993) Role of inoculum preparation and density on the bioremediation of 2,4-d-contaminated soil by bioaugmentation. Appl Microbiol Biotechnol 38: 681–687. Corke CT, NJ Bunce, A-L Beaumont, RL Merrick (1979) Diazonium cations as intermediates in the microbial transformations of chloroanilines to chlorinated biphenyls, azo compounds and triazenes. J Agric Food Chem 27: 644–646. D’Annibale A, F Rosetto, F Federici, M Petruccioli (2006) Role of autochthonous filamentous fungi in bioremediation of a soil historically contaminated with aromatic hydrocarbons. Appl Environ Microbiol 72: 28–36. Estrella MR, ML Brusseau, RS Maier, IL Pepper, PJ Wierenga, RM Miller (1993) Biodegradation, sorption, and transport of 2,4-dichlorophenoxyacetic acid in saturated and unsaturated soils. Appl Environ Microbiol 59: 4266–4273. Garrison AW, P Schmitt, O Martens, A Kettrup (1996) Enantiomeric selectivity in the environmental degradation of dichloroprop as determined by high-performance capillary electrophoresis. Environ Sci Technol 30: 2449–2455. Gintautas PA, SR Daniel, DL Macalady (1992) Phenoxyalkanoic acid herbicides in municipal landfill leachates. Environ Sci Technol 26: 517–521. Greer LE, DR Shelton (1992) Effect of inoculant strain and organic matter content on kinetics of 2,4-dichlorophenoxyacetic acid degradation in soil. Appl Environ Microbiol 58: 1459–1465. Hammill TB, RL Crawford (1996) Degradation of 2-sec-butyl-4,6-dinitrophenol (Dinoseb) by Clostridium bifermentans KMR-1. Appl Environ Microbiol 62: 1842–1846. Kaake RH, DJ Roberts, TO Stevens, RL Crawford, DL Crawford (1992) Bioremediation of soils contaminated with the herbicide 2-sec-butyl-4,6-dinitrophenol (Dinoseb). Appl Environ Microbiol 58: 1683–1689. Kjaer J, P Olsen, T Henriksen, M Ullum (2005) Leaching of metribuzin metabolites and the associated contamination of a sandy Danish aquifer. Environ Sci Technol 39: 8374–8381. Lee JK, KS Kyung (1995) Uptake of 3,3′,4,4′-tetrachloroazobenzene soil residues by rice (Oryza sativa L.). J Agric Food Chem 43: 519–523. Martin M, E Ferrer, R Alonso, J Fernández (1995) Bioremediation of soil contaminated by propachlor using native bacteria. Int Biodet Biodeg 213–225. Ogram AV, RE Jessup, LT Ou, PSC Rao (1985) Effects of sorption on biological degradation rates of 2,4dichlorophenoxyacetic acid in soils. Appl Environ Microbiol 49: 582–587. Pieper DH, R Winkler, H Sandermann (1992) Formation of a toxic dimerization product of 3,4-dichloroaniline by lignin peroxidase from Phanerochaete chrysosporium. Angewandte Chemie 104: 60–61. Roberts DJ, RH Kaake, SB Funk, DL Crawford, RL Crawford (1993) Anaerobic remediation of dinoseb from contaminated soil. An on-site demonstration. Applied Biochemistry and Biotechnology – Part A Enzyme Engineering and Biotechnology 39–40: 781–789. Sandermann H Jr., W Heller, N Hertkorn, E Hoque, D Pieper, R Winkler (1998) A new intermediate in the mineralization of 3,4-dichloroaniline by the white rot fungus Phanerochaete chrysosporium. Appl Environ Microbiol 64: 3305–3312. Silva E, AM Fialho, I Sa-Correia, RG Burns, LJ Shaw (2004) Combined bioaugmentation and biostimulation to cleanup soil contaminated with high concentrations of atrazine Environ Sci Technol 38: 632–637. Simmons KE, RD Minard, J-M Bollag (1987) Oligomerization of 4-chloroaniline by oxidoreductases. Environ Sci Technol 21: 999–1003. Tas DO, SG Pavlostathis (2005) Microbial reductive transformation of pentachloronitrobenzene under methanogenic conditions. Environ Sci Technol 39: 8264–8272.

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Tatsumi K, A Freyer, RD Minard, J-M Bollag (1994) Enzyme-mediated coupling of 3,4-dichloroaniline and ferulic acid: a model for pollutant binding to humic materials. Environ Sci Technol 28: 210–215. Tett VA, AJ Willetts, HM Lappin-Scott (1994) Enantioselective degradation of the herbicide mecoprop [2-2methyl-4-chlorophenoxypropionic acid] by mixed and pure bacterial cultures. FEMS Microbiol Ecol 14: 191–200. You IS, R Bartha (1982) Stimulation of 3,4-dichloroaniline mineralization by aniline. Appl Environ Microbiol 44: 678–681. Zipper C, K Nickel, W Angst, H-PE Kohler (1996) Complete microbial degradation of both enantiomers of the chiral herbicide mecoprop [2-4-chloro-2-methylphenoxypropionic acid] in an enantioselective manner by Sphingomonas herbicidovorans sp. nov. Appl Environ Microbiol 62: 4318–4322.

PART 6 MILITARY WASTE EXPLOSIVES A wide range of nitro compounds has been used as military explosives, including nitrotoluenes, nitrate esters, and nitramines. There is concern over the destruction of these toxic and potentially explosive products, and considerable effort has been devoted to the development of microbiological processes of destruction (Kaplan 1990; Crawford 1995). Details of the degradation pathways of the relevant compounds have been given (Walker and Kaplan 1992), and general comments on the degradation of aromatic nitro compounds may be found in a review (Spain 1995a) and in a book (Spain 1995b). A review specifically devoted to 2,4,6-trinitrotolune is also available (Esteve-Núnez et al. 2001). Although the biodegradation of aromatic nitro compounds has been discussed in Chapter 9, Part 5, attention is drawn to the role of fungi (Bayman and Radkar 1997) including Phanerochaete chrysosporium (Bumpus and Tatarko 1994), and the significance of association between reduced metabolites and soil organic matter (Achnich et al. 1999; Thorn and Kennedy 2002). Compared with the wastes that have already been considered, these compounds like triazines are distinguished by a higher N/C ratio, so that although they may serve as sources of both carbon and nitrogen, when only the nitrogen is used by some organisms an additional carbon source must be added. The biodegradation of aromatic nitro compounds has already been discussed in Chapter 9, Part 5.

NITROARENES Only limited data are available on full- or pilot-scale operations, and the results from these may present an oversimplified view. In general reduction to amines is the initial—and sometimes the only—transformation. 1. Indigenous microorganisms from a munitions-contaminated site reduced only a single nitro group of 2,4- and 2,6-dinitrotoluene (DNT), and 2,4,6-trinitrotoluene (TNT) to amines, while TNT was mineralized to ca. 10% (Bradley et al. 1994). A preliminary experiment using soil contaminated with both TNT and nitramines was carried out by adding phosphate and starch to the surface. The anaerobic flora that developed reduced the nitro groups of TNT, and subsequently the reduced compounds were removed though their fate was not established (Funk et al. 1993). One important additional observation was that under aerobic conditions or under anaerobic conditions at alkaline pH, polymerization of the intermediate amines took place. This experiment has since been extended to the evaluation of an in situ bioreactor (Funk et al. 1995). 2. Bioreactors containing an undefined anaerobic consortium reduced TNT to 2,4,6triaminotoluene (TAT) in the presence of glucose (Daun et al. 1998). The sorption of TAT to montmorillonite clay was irreversible, and the substrate could not be released by solvent extraction or by acid or alkaline treatment. Similar results were obtained with humic acids in which covalent reactions with carbonyl or activated C=C bonding presumably occurred. Results from laboratory experiments using 14C-labeled TNT in reactors to which

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molasses was added as carbon source, showed that after 9 weeks 83% of the radioactivity was recovered in soil components (humin, humic acids, and fulvic acids) (Drzyzga et al. 1998). In a full-scale application where sucrose was added to obtain anaerobic conditions, reduction of the nitro groups to amines was observed, followed by irreversible binding to the soil. Subsequent aerobic treatment was used to remove the sucrose and its fermentation products (Lenke et al. 1998). These results illustrate the cardinal issue of the association of metabolic products from TNT with soil components, and this in turn may substantially reduce the effectiveness of remediation. 3. Simultaneous degradation of 2,4- and 2,6-DNT has been achieved in a fluidized-bed biofilm reactor (Lendemann et al. 1998) that was successfully operated with contaminated groundwater containing 2- and 4-nitrotoluenes and 2,4- and 2,6-DNTs. The nitrite that was produced during degradation was recovered as nitrate. 4. Partial reduction by clostridia has been demonstrated and has already been discussed in Chapter 9, Part 5. The effectiveness of clostridia in the anaerobic digestion of a number of nitroaromatics has prompted development of a procedure for the production of spores of Clostridium bifermentans, and a medium for their effective production has been developed (Sembries and Crawford 1997).

NITRODIPHENYLAMINES Although these are no longer used as explosives, they are formed from diphenylamine that is used as a stabilizer for nitrocellulose explosives. The metabolism of nitrodiphenylamines has been examined under anaerobic conditions, and phenazine and 4-aminoacridine that are cyclization products of the initially formed 2-aminodiphenylamine have been identified (Drzyzga et al. 1996).

NITRATE ESTERS A review (White and Snape 1993) has been devoted to this topic, and only very brief comments need be added. The reductive biotransformation of glycerol trinitrate by strains of Bacillus thuringiensis/ cereus or Enterobacter agglomerans involves the expected successive loss of nitrite with formation of glycerol (Meng et al. 1995). The transformation of pentaerythritol tetranitrate by Enterobacter cloacae that used this as a source of nitrogen proceeded analogously, with further metabolism of the two hydroxymethyl groups to the dialdehyde (Binks et al. 1996). The enzymes that produce nitrite from the nitrate esters are reductases rather than hydrolases, and the enzyme from E. cloacae is strongly inhibited by steroids, and is capable of the reduction of cyclohex-2-ene-1-one (French et al. 1996). In a medium containing glucose and ammonium nitrate, glyceryl trinitrate is transformed by Penicillium corylophilum to the dinitrate and the mononitrate before complete degradation (Zhang et al. 1997.). In contrast, the metabolism of glyceryl trinitrate by Phanerochaete chrysosporium involves the production of the physiologically important nitric oxide (Servent et al. 1991).

NITRAMINES Of the nitramines used as explosives, greatest attention has been directed to the degradation of 1,3,5-trinitro-hexahydro-1,3,5-triazine (RDX) and octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine (HMX). RDX is degraded by several species of Enterobacteriaceae (Kitts et al. 1994), and is used by Stenotrophomonas sp. strain PB1 as a source of nitrogen with the production of metabolites formed by loss of two nitro groups and a single nitrogen atom from the triazine. Their structures were tentatively put forward (Binks et al. 1995), although an alternative anaerobic pathway involving reduction of the nitro groups to nitroso and a series of transformations yielding hydrazine, dimethylamine, and 1,2-dimethylhydrazine has been presented (McCormick et al. 1981). In addition to military uses, it should be noted that aromatic nitro compounds such as trifluralin (2,6-dinitro-N,N-dipropyl-4-(trifluoromethyl)aniline and dinoseb (2,4-dinitro-6-sec-butylphenol)

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have been used as herbicides, and parathion (O,O-diethyl-O-(4-nitrophenyl)phosphorothioate) as an insecticide. In addition, aromatic nitro compounds are of considerable industrial value for the production of the anilines (Hartter 1985) that are components of several agrochemicals.

CONCLUSIONS Although the possibility of bioremediation of sites contaminated with nitrotoluene waste is clearly possible, important issues should be clearly appreciated. These include (a) additional carbon sources may be necessary to accomplish partial or complete reduction of nitro groups, (b) under aerobic conditions, dimeric azo compounds may be formed as terminal metabolites, and (c) aromatic amines may be incorporated into humic material by covalent bonding and thereby resist further degradation.

CHEMICAL WARFARE AGENTS Brief notes are added on phosphorofluoridates even though their destruction by microbial activity— though clearly possible—is limited by their toxicity to the requisite microorganisms. One of the motivations for their inclusion is the fact that the hydrolytic enzyme(s) responsible for defluorination—organophosphorus acid anhydrase (OPA)—is widespread, and is found in a number of bacteria (Landis and DeFrank 1990). The microbial hydrolysis of organophosphorus pesticides and cholinesterase inhibitors is accomplished by several distinct enzymes, which are collectively termed organophosphorus acid anhydrases (OPAs). These have been reviewed (DeFrank 1991), so that only a few additional comments are necessary. The enzymes from Pseudomonas putida strain MG and Flavobacterium sp. (ATCC 27551) are carried on plasmids (Mulbry et al. 1986) and are membrane-bound. Although they hydrolyze a range of phosphorothioate esters at rates greatly exceeding those for chemical hydrolysis and have been proposed for application to destruction of these (Munnecke 1976), they hydrolyze phosphorofluoridates much more slowly (Dumas et al. 1989). This, together with the fact that they are membranebound makes them of somewhat limited attraction for the destruction of phosphorofluoridates. Two other enzymes have attracted attention for the hydrolysis of phosphorofluoridates: one from Bacillus stearothermophilus is also membrane-bound, whereas the other from Alteromonas sp. strain JD 6.5 is a soluble enzyme. OPA activity has also been observed in a number of other organisms (DeFrank et al. 1993). The enzyme designated OPA-2 has been purified from Alteromonas strain JD 6.5, and has a molecular mass of 60 kDa (DeFrank and Cheng 1991). There has been considerable speculation on the natural substrate for OPA activity, and a surprising fact emerged from a detailed comparison of the amino acid sequence. This revealed a high degree of homology with an aminopeptidase from Escherichia coli and that the OPA anhydrase had high prolidase activity (Cheng et al. 1996). All these observations underscore the potential for application of appropriate OPAs to the destruction of organophosphorus compounds with anticholinesterase activity (Cheng and Calomiris 1996). However, since, hydrolysis results in release of fluoride, the possibility of its subsequent incorporation into organic substrates to produce fluoroacetate and 4-fluorothreonine (Reid et al. 1995) may be worth consideration.

REFERENCES Achnich A, E Fernandes, J-M Bollag, H-J Knackmuss, H Lenke (1999) Covalent binding of reduced metabolites of [15N3]TNT to soil organic matter during a bioremediation process analyzed by 15N NMR spectroscopy. Environ Sci Technol 33: 4448–4456. Bayman P, GV Radkar (1997) Transformation and tolerance of TNT (2,4,6-trinitrotoluene) by fungi. Int Biodet Biorem 39: 45–53. Binks PR, CE French, S Nicklin, NC Bruce (1996) Degradation of pentaerythritol tetranitrate by Enterobacter cloacae PB2. Appl Environ Microbiol 62: 1214–1219.

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Binks PR, S Nicklin, NC Bruce (1995) Degradation of hexahydro-1,3,5-trinitro-1,3,5-triazine RDX by Stenotrophomonas maltophilia. PB1 Appl Environ Microbiol 61: 1318–1322. Bradley PM, FH Chapelle, JE Landmeyer, JG Schumacher (1994) Microbial transformation of nitroaromatics in surface soils and aquifer materials. Appl Environ Microbiol 60: 2170–2175. Bumpus JA, M Tatarko (1994) Biodegradation of 2,4,6-trinitrotoluene by Phanerochaete chrysosporium: identification of initial degradation products and the discovery of a metabolite that inhibits lignin peroxidases. Curr Microbiol 28: 185–190. Cheng T-C, JJ Calomiris (1996) A cloned bacterial enzyme for nerve agent decontamination. Enz Microbiol Technol 18: 597–601. Cheng T-C, SP Harvey, GL Chen (1996) Cloning and expression of a gene encoding a bacterial enzyme for decontamination of organophosphorus nerve agents and nucleotide sequence of the enzyme. Appl Environ Microbiol 62: 1636–1641. Crawford RL (1995) The microbiology and treatment of nitroaromatic compounds. Curr Opinion Biotechnol 6: 329–336. Daun G, H Lenke, M Reuss, H-J Knackmuss (1998) Biological trearment of TNT-contaminated soil. 1. Anaerobic cometabolic reduction and interaction of TNT and metabolites with soil components. Environ Sci Technol 32: 1956–1963. DeFrank JJ (1991) Organophosphorus cholinesterase inhibitors: detoxification by microbial enzymes. In Applications of Enzyme Biotechnology (Ed JW Kelly and TO Baldwin), pp. 165–180. Plenum Press, New York. DeFrank JJ, T-C Cheng (1991) Purification and properties of an organophosphorus acid anhydrase from a halophilic bacterial isolate. J Bacteriol 173: 1938–1943. DeFrank JJ, WT Beaudry, T-C Cheng, SP Harvey, AN Stroup, LL Szafraniec (1993) Screening of halophiliuc bacteria and Alteromonas species for organophosphorus hydrolyzing enzyme activity. Chem-Biol Interact 87: 141–148. Drzyzga O, A Schmidt, K-H Blotevogel (1996) Cometabolic transformation and cleavage of nitrodiphenylamines by three newly isolated sulfate-reducing bacterial strains. Appl Environ Microbiol 62: 1710–1716. Drzyzga O, D Bruns-Nagel, T Gorontzy, K-H Blotevogel, E von Löw (1998) Incorporation of 14C-labeled 2,4,6-trinitrotoluene metabolites into different soil fractions after anaerobic and anaerobic-aerobic treatment of soil/molasses mixtures. Environ Sci Technol 32: 3529–3535. Dumas DP, JR Wild, FM Raushel (1989) Diisopropylfluorophosphate hydrolysis by an organophosphate anhydrase from Pseudomonas diminuta. Biotechnol Appl Biochem 11: 235–243. Esteve-Núnez A, A Caballero, JL Ramos (2001) Biological degradation of 2,4,6-trinitrotoluene. Microbiol Mol Biol Rev 65: 335–352. French CE, S Nicklin, NC Bruce (1996) Sequence and properties of pentaerythritol tetranitrate reductase from Enterobacter cloacae PB2. J Bacteriol 178: 6623–6627. Funk SB, DJ Roberts, DL Crawford, RL Crawford (1993) Initial-phase optimization for bioremediation of munition compound-contaminated soils. Appl Environ Microbiol 59: 2171–2177. Funk SB, DL Crawford, RL Crawford, G Mead, W Davis-Hoover (1995) Full-scale anaearobic bioremediation of trinitrotoluene TNT contaminated soil. Appl Biochem Biotechnol 51/52: 625–633. Hartter DR (1985) The use and importance of nitroaromatic compounds in the chemical industry. In Toxicity of Nitroaromatic Compounds (Ed D Rickert), pp. 1–13. Hemisphere Publishing Corporation, Washington, DC. Kaplan DL (1990) Biotransformation pathways of hazardous energetic organo-nitro compounds. In Biotechnology and Biodegradation (Eds D Kamely, A Chakrabarty and GS Omenn), Vol. 4, pp. 155–181. Gulf Publishing Company, Houston. Kitts CL, DP Cunningham, PJ Unkefer (1994) Isolation of three hexahydro-1,3,5-trinitro-1,3,5-triazinedegrading species of the family Enterobacteriaceae from nitramine explosive-contaminated soil. Appl Environ Microbiol 60: 4608–4711. Landis WG, JJ DeFrank (1990) Enzymatic hydrolysis of toxic organofluorophosphate compounds. In Biotechnology and Biodegradation (Eds D Kamely, A Chakrabarty, GS Omenn), Vol. 4, pp. 183–201. Gulf Publishing Company, Houston, USA. Lendemann U, JC Spain, BF Smets (1998) Simultaneous biodegradation of 2,4-dinitrotoluene and 2,6-dinitrotoluene in an anerobic fluidized-bed biofilm reactor. Environ Sci Technol 32: 82–87. Lenke H, J Warrelmann, G Daun, K Hund, U Sieglen, U Walter, H-J Knackmuss (1998) Biological treatment of TNT-contaminated soil. 2. Biologically induced immobilization of the contaminants and full-scale application. Environ Sci Technol 32: 1964–1971.

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McCormick NG, JH Cornell, AM Kaplan (1981) Biodegradation of hexahydro-1,3,5-trinitro-1,3,5-triazine. Appl Environ Microbiol 42: 817–823. Meng M, W-Q Sun, LA Geelhaar, G Kumar, AR Patel, GF Payne, MK Speedie, JR Stacy (1995) Denitration of glycerol trinitrate by resting cells and cell extracts of Bacillus thuringiensis/cereus and Enterobacter agglomerans. Appl Environ Microbiol 61: 2548–2553. Microbiol Rev 45: 180–209. Mulbry WW, JS Karns, PC Kearney, JO Nelson, CS McDaniel, JR Wild (1986) Identification of a plasmidborne parathion hydrolase gene from Flavobacterium sp. by Southern hybridization with opd from Pseudomonas diminuta. Appl Environ Microbiol 51: 926–930. Munnecke DM (1976) Enzymatic hydrolysis of organophosphate insecticides, a possible pesticide disposal method. Appl Environ Microbiol 32: 7–13. Preuss A, J Fimpel, G Diekert (1993) Anaerobic transformation of 2,4,6-trinitrotoluene (TNT). Arch Microbiol 159: 345–353. Reid KA, JTG Hamilton, RD Bowden, D O’Hagan, L Dasaradhi, MR Amin, DB Harper (1995) Biosynthesis of fluorinated secondary metabolites by Streptomyces cattleya. Microbiology (UK) 141: 1385–1393. Sembries S, RL Crawford (1997) Production of Clostridium bifermentans spores as inoculum for bioremediation of nitroaromatic contaminants. Appl Environ Microbiol 63: 2100–2104. Servent D, C Ducrorq, Y Henry, A Guissani, M Lenfant (1991) Nitroglycerin metabolism by Phanerochaete chrysosporium: evidence for nitric oxide and nitrite formation. Biochim biophys Acta 1074: 320–325. Spain JC (1995a) Biodegradation of nitroaromatic compounds. Annu Rev Microbiol 49: 523–555. Spain JC (Ed) (1995b) Biodegradation of Nitroaromatic Compounds. Plenum Press, New York. Thorn KA, KR Kennedy (2002) 15N NMR investigation of the covalent binding of reduced TNT amines to soil humic acid, model compounds, and lignocellulose. Environ Sci Technol 36: 3787–3796. Walker JE, DL Kaplan (1992) Biological degradation of explosives and chemical agents. Biodegradation 3: 369–385. White GF, JR Snape (1993) Microbial cleavage of nitrate esters, defusing the environment. J Gen Microbiol 139: 1947–1957. Zhang YZ, ST Sundaram, A Sharma, BW Brodman (1997) Biodegradation of glyceryl trinitrate by Penicillium corylophilum Dierckx. Appl Environ Microbiol 63: 1712–1714.

PART 7 GROUNDWATER CONTAMINATION BENZENE/TOLUENE/ETHYLBENZENE/XYLENES These compounds along with trimethylbenzenes are components of automobile gasoline, and may enter aquatic systems and eventually groundwater through spillage at filling stations, leakage from underground storage tanks, or during transport. Aerobic Conditions All these substrates can be degraded under aerobic conditions and, although there appear to be important differences among the xylene isomers and mutant strains have been isolated that can degrade all three isomers (Di Lecce et al. 1997). Reviews have covered various aspects of this problem: 1. Biodegradation in gas-condensate-contaminated groundwater with emphasis of the role of inorganic nutrients and oxygen (Morgan et al. 1993) 2. Determinative factors for treating aquifer plumes after accidental spillage of automobile fuel with discussions on the design of systems, and the role of the indigenous microflora (Salanitro 1993) 3. General procedures including the provision of oxygen and comments on anaerobic treatment (Bowlen and Kosson 1995) The pathways for the aerobic biodegradation of alkylated benzenes have been elucidated in extensive investigations and have been discussed in Chapter 8, Part 1, so that only salient features are briefly summarized here. The genes for the degradation of toluene may be either chromosomal or

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plasmid-borne. The latter may, therefore be lost in the absence of selection pressure. The degradation of toluene may occur by different pathways: 1. Dioxygenation or successive monooxygenation of the substrate, with subsequent ring fission leaving the alkyl group intact 2. Oxidation of the alkyl group to a carboxylic acid before ring fission. For xylenes, it appears that there are significant differences in the degradability of the isomers, o-xylene being apparently more recalcitrant The results of laboratory experiments with fluidized-bed reactors treating toluene-contaminated influent (Massol-Deyá et al. 1997) came to important conclusions: 1. Toluene-degrading strains in the biofilm were displaced by groundwater organisms from a putatively pristine source. 2. These organisms were able to establish a stable microbial community and were able to bring about a decrease in toluene concentration in the effluent from 0.140 to 0.063 mg/L. These results clearly show the advantage and desirability of using indigenous organisms under appropriate conditions, and that they may effectively degrade relatively low substrate concentrations. The latter is consistent with the ability of bacteria in natural aquatic systems to utilize low substrate concentrations, which has been noted in Chapter 4. An experiment on toluene biodegradation under field and laboratory conditions provided results of value in the design of laboratory simulation experiments, and illustrated the caution required in assessing the fate of contaminants. Experiments were carried out under three conditions: (i) in flow-through horizontal columns containing sediment and rocks, (ii) in shaking cultures containing sediments, rocks, or plant material, and (iii) in situ in a contaminated stream (Cohen et al. 1995). There were a number of important conclusions: a. The rates of biodegradation observed in the shake-flask experiment more closely paralleled those from the in situ experiment than those using the flow-through column system b. Biodegradation took place on streambed surfaces particularly on rocks, so that the formation of biofilms was essential c. Conversion into CO2 was incomplete, and 15% of the substrate was converted into soluble metabolites, and 62% to biomass or insoluble material Consistent with this, it was observed that, during the metabolism of 14C-labeled BTEX although the substrates were mineralized to ca. 70%, ca. 20% of the label in toluene and ca. 30% in o-xylene were found in humus. It was suggested that the alkylated catechol metabolites might plausibly be responsible for this association (Tsao et al. 1998). The effect of increased oxygen was examined in contaminated groundwater (Gibson et al. 1998). Oxygen was supplied by diffusion from silicone tubing to areas within the plume that were oxygen-deficient due to the activity of indigenous bacteria. The rate of aerobic BTEX degradation increased at the expense of anaerobic degradation, and resulted in the degradation of both benzene and ethylbenzene that are generally less readily degraded under anaerobic conditions. Anoxic or Anaerobic Conditions In many situations, such as in deep groundwater, oxygen concentration may be severely limiting due to its consumption by surficial aerobic organisms, or its low rate of transport into the system. Therefore, there has been great interest in the anaerobic degradation of BTEX. Although this has been

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demonstrated for benzene under anaerobic conditions by methanogenic consortia, sulfate-reducing bacteria, and Fe(III)-reducing bacteria, greatest attention has been directed in laboratory studies to the anaerobic degradation of toluene in the presence of nitrate. This is supported by the fact that the appropriate organisms are widely distributed (Fries et al. 1994). Some of the important principles may be illustrated from two groups of experiments. The first used a laboratory microcosm and a field experiment with a contaminated aquifer (Barbaro et al. 1992). Although the results clearly showed the effectiveness of adding nitrate, there were significant differences in the extent to which the substrates were degraded: for example, benzene appeared recalcitrant, and the xylenes and ethyl benzene were less degradable than toluene. Nitrate was apparently consumed by indigenous organisms at the expense of other carbon substrates, so that this was effectively removed from the system. In this experiment—and possibly in general—sorption of these substrates to organic carbon was not a limiting factor. In the second example, the effect of nitrate or sulfate additions was examined by adding either of these, together with samples of BTEX to a contaminated aquifer in the form of “slugs” of purified groundwater (Reinhard et al. 1998). For nitrate additions, the rates of removal of the substrates were in the order benzene < o-xylene < ethylbenzene < toluene = m-xylene, the last three being removed within 6 d. For sulfate additions, removal of toluene and all the xylenes occurred within 50 d. Metabolites characteristic of degradation (Beller et al. 1995) were isolated as confirmation of biotransformation. In the interpretation of such results, several important considerations should be taken into account in the application of this procedure. a. Although the results (Beller et al. 1995) were broadly in agreement with those from laboratory microcosm experiments, the rates of degradation of toluene and m- and p-xylenes seemed to be lower. Benzene and o-xylene seem to be considerable more recalcitrant than other components of BTEX. b. In an aquifer that was contaminated with BTEX, it has been shown that the methylbenzylsuccinates arising from xylenes were present, although benzylsuccinate, which would be expected from the toluene that was present was not observed. This was presumed to have undergone further degradation in the aquifer (Beller 2002). The application of deuterated toluene in assessing anaerobic biodegradation (Fischer et al. 2006) has already been noted in Chapter 6, Part 1. A review (Lovley 1997) has summarized the various strategies and suggests that uncertainties, particularly in the bioremediation of benzene, can only be resolved by greater emphasis on field-oriented studies, and a better understanding of the reactions involved and the factors that limit the rates of degradation. These experiments should be viewed against the results of extensive laboratory investigations of the anaerobic degradation and transformation of the individual components of BTEX. The degradation of ethylbenzene and n-propylbenzene—though not xylenes—has been shown in pure cultures of denitrifying bacteria (Rabus and Widdel 1995). In crude oil, toluene, ethylbenzene, and m-xylene were degraded, though neither o-xylene nor p-xylene (Rabus and Widdel 1996). Degradation of toluene by the sulfate-reducing anaerobe Desulfobacula toluolica has been shown (Rabus et al. 1993), and by another sulfate-reducing bacterium with a novel 16S rRNA gene sequence (Beller et al. 1996). The mechanism of these anaerobic degradations is complex and reviews have been given by some of the pioneers in this area (Evans and Fuchs 1988; Fuchs et al. 1994). It is important to note that benzylsuccinates and benzylfumarates have been proved to be intermediates during the oxidation of alkylbenzenes to carboxylates (Biegert et al. 1997; Beller and Spormann 1997). As noted above, these compounds together with the analogous compounds that would be produced from xylenes have been used in support of the active in situ anaerobic degradation of toluene and xylenes (Beller et al. 1995, Beller 2002). In addition, the potential for application of the analogous carboxylates of naphthalene, benzothiophene, and benzofuran has been suggested (Safinowski et al. 2006).

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HALOGENATED ALKANES AND ALKENES Tetrachloromethane The degradation of tetrachloromethane by Pseudomonas stutzeri strain KC involves hydrolysis to CO2 by a mechanism involving the naturally produced pyridine-2,6-dithiocarboxylic acid (Lewis et al. 2001): details have already been discussed in Chapter 7, Part 3. This organism was used in field evaluation at a site at which the indigenous flora was ineffective, and acetate was used as electron donor (Dybas et al. 2002). One novel feature was inoculation at a series of wells perpendicular to the established flow of the groundwater plume. Effective removal of tetrachloromethane was sustained over a period of 4 years, and transient levels of chloroform and H2S disappeared after reducing the concentration of acetate. Chloroethenes Tetrachloroethene (TeCE; perchloroethene [PCE]) and trichloroethene (TCE) have been used extensively as solvents and degreasing agents, and substantial effect has been devoted to the bioremediation of contaminated groundwater. Determinative issues are the relatively low substrate concentrations combined with the large volumes of groundwater that are involved. Remediation is made more difficult by the high density and poor water solubility of chlorinated ethenes, and the possibility of improvement by using cyclodextrin to enhance solubility has been examined for a TeCE-contaminated aquifer (Tick et al. 2003). There has therefore been greater interest in the application of in situ procedures than in more conventional “pump and treat” on-site methods. To appreciate the microbiology of the procedures, a brief summary of the relevant aspects is given. Reviews provide additional details (Ensley 1991; Arp 1995), and relevant details of the reactions have already been given. A valuable review (Lee et al. 1998) includes a discussion of microbial processes, and provides illustrative examples of both in situ and natural attenuation. An overview of possible procedures has been given (Stroo et al. 2003). Aerobic Conditions TCE is not able to support the growth of a single organism, but it is susceptible to cooxidation by oxygenases elaborated by organisms during growth with structurally unrelated substrates. A review of methanotrophic bacteria (Hanson and Hanson 1996) contains a useful account of their application to bioremediation of TCE-contaminated sites. There are several bacterial oxygenases that can bring about cooxidation of TCE after induction with the relevant substrates: methane monooxygenase in Methylosinus trichosporium strain OB3b (Jahng and Wood 1994), propane monooxygenase in Mycobacterium vaccae (Vanderberg et al. 1995), and ammonia monooxygenase (Arciero et al. 1989; Vannelli et al. 1990). Although the methane monooxygenase system involves an epoxide intermediate that may form CO and toxifies the organism, this may be overcome by addition of a suitable reductant such as formate (Henry and Grbic-Galic 1991). An alternative strategy is the intermittent provision of substrate TCE (Walter et al. 1997). Pseudomonas putida strain F1 elaborates a dioxygenase for the degradation of toluene, and this is able to cooxidize TCE that may also be carried out by growth with phenol (Nelson et al. 1988). A similar situation prevails for the isopropylbenzene-degrading Rhodococcus erythropolis strain BD-2 (Dabrock et al. 1994). Increased biodegradation of TCE has also been observed in a constructed strain of Escherichia coli carrying the genes for the large subunit of toluene terminal oxygenase (from Pseudomonas putida F1), the small subunit of biphenyl terminal oxygenase, and the ferredoxin and ferredoxin reductase (from P. pseudoalcaligenes KF707) (Furukawa et al. 1994). Monooxygenase systems may also be involved in the degradation of toluene and bring about the cooxidation of TCE under suitable growth conditions. P. cepacia strain G4 carries out ortho hydroxylation,

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P. mendocina KR-1 carries out para hydroxylation, and Ralstonia (Pseudomonas) pickettii carries out para hydroxylation, although it was originally reported to carry out meta hydroxylation. The complexity introduced by exposure of an established mixed culture growing with a single substrate to an alternative cosubstrate is illustrated by the following. A stable mixed culture of Pseudomonas putida mt-2, P. putida F1, P. putida GJ31, and Burkholderia cepacia G4 growing with limited concentrations of toluene was established. Exposure to TCE for a month resulted in the loss of viability of the last three organisms, and resulted in a culture dominated by P. putida mt-2 from which mutants had fortuitously arisen (Mars et al. 1998). Two different types of experiments on bioremediation of sites contaminated with TCE have been carried out, and have been preceded by, and taken advantage of the valuable results obtained in microcosm experiments. Application of Indigenous Bacteria One procedure used an indigenous microflora enriched by the addition of substrates that induced the appropriate oxygenases. Microcosm experiments (Hopkins et al. 1993b) demonstrated the differential effects of phenol, toluene, methane, and ammonia. Whereas the first two substrates were equally effective in removing both TCE and trans-1,2-dichloroethene (DCE), methane was only marginally effective in removing TCE though more effective in removing trans-1,2-DCE, while ammonia was less effective in removing all the substrates. It was shown that in bioreactors fed with phenol under various conditions, maximum removal of TCE occurred when phenol was supplied as a pulse every 24 h. In this reactor there was a greater diversity of microorganisms and higher numbers that degraded phenol and TCE (Shih et al. 1996). Field evaluation of this procedure has been carried out, using phenol or toluene or methane as oxygenase inducers (Hopkins et al. 1993a; Hopkins and Mccarty 1995). The effectiveness of the various substrates for removal of a number of chloroethenes is given in Table 14.3. Using phenol or toluene, therefore there remains no serious concern over the recalcitrance of chloroethenes—and particularly of vinyl chloride produced by partial anaerobic dechlorination of TCE. In this experiment, the concentrations of residual phenol and toluene were below the standards set for drinking water. A full-scale treatment of TCE-contaminated groundwater used pulsed addition of toluene and supplied oxygen both as O2 and H2O2: a high level of TCE removal was achieved over a 410-d period (McCarty et al. 1998). In a further development, in situ stripping of TCE was used to release the TCE from the dense water phase, and this was then mixed with toluene, hydrogen peroxide, oxygen, and a tracer and was pumped down-flow into the aquifer where degradation occurred. A high level of removal was accomplished with the stripping, and a further amount by biodegradation that resulted in an effluent with the concentration of TCE reduced to 250 µg/L (Goltz et al. 2005).

TABLE 14.3 Influence of Growth Substrates on the Removal of Chloroethenes % Removal of Substrate Growth Substrate

Concentration (mg/L)

TCE

1,1-DCE

cis-DCE

trans-DCE

MCE

Methane 6.6 19 — 43 90 95 Phenol 12.5 94 54 92 73 >98 Toluene 9 93 — >98 75 — Note: TCE = trichloroethene; 1,1-DCE = 1,1-dichloroethene; cis-DCE = cis-1,2-dichloroethene; trans-DCE = trans-1,2dichloroethene; MCE = chloroethene (vinyl chloride)

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Microbiological investigations after treatment have revealed several important additional facts: a. The adverse effect of 1,1-DCE on the populations was restored after its removal, and the original population diversity was recovered only after treatment with toluene and TCE. The dominant Gram-negative taxa belonged to the genera Comamonas-Variovorax, Azoarcus, and Burkholderia (Fries et al. 1997a); b. Although many Gram-negative strains hybridized to a toluene 2-monooxygenase probe, not all positive strains effectively degraded TCE, and none were as effective as Burkholderia cepacia strain G4. Since many indigenous strains were able to degrade toluene—though not all of them were able to metabolize TCE—such strains would dominate populations in the absence of TCE (Fries et al. 1997b). This is consistent with the results of a laboratory study using B. cepacia strain G4 (Mars et al. 1996). Application of Exogenous Bacteria In the second procedure, organisms with an established degradation potential for TCE were injected at the site. 1. Cells of Methylosinus trichosporium strain OB3b were added and became attached to the sediments to form an “in situ bioreactor.” The concentration of TCE in samples withdrawn through the biofilter region fell rapidly, although after 40 d the concentrations had reverted to background levels (Duba et al. 1996). The major remaining unresolved issue was therefore how the degradative activity could be maintained during the length of time required for effective bioremediation. 2. In the light of possible concern over the introduction of phenol into aquifers, mutant strains of Burkholderia (Pseudomonas) cepacia strain G4 in which toluene 2-monooxygenase is constitutive have been examined in microcosm experiments. Effective removal of TCE could be demonstrated (Krumme et al. 1993; Munakata-Marr et al. 1996), although in the first experiment serious loss in the viability of the cells occurred during the 10-week incubation. This aspect therefore presents a remaining challenge. Experiments have been conducted in a contaminated aquifer to examine the effect of Triton X-100 on increasing the release of TCE from sediments (Sahoo et al. 1998). The increase was rather modest (ca. 30%), and is consistent with the limited effect of surfactants in increasing the availability of PAHs, which has been discussed in Part 2 of this chapter. Anoxic and Anaerobic Conditions The anaerobic dechlorination of chloroethenes has received considerable attention, and has been encouraged by concern that partial dechlorination could result in the production of vinyl chloride as a terminal metabolite. After induction by growth of Desulfomonile tiedjei DCB-1 with 3-chlorobenzoate, partial dechlorination of TeCE to TCE and cis-1,1-DCE has been observed (Cole et al. 1995; Townsend and Suflita 1996). Dehalogenases for both TCE and TeCE have been examined in a strain of Sulfurospirillum (Dehalospirillum) multivorans (Neumann et al. 1994), and it has been shown that the dechlorination of TeCE in anaerobic aquifer slurries was stimulated by C-3 carboxylic acids or ethanol (Gibson and Sewell 1992). Particular attention has centered on the application of organisms such as Dehalococcoides ethenogenes that can carry out complete dechlorination to ethene. Details of the anaerobic degradation of chloroalkanes and chloroalkenes have been discussed in detail in Chapter 7, Part 3, and the application of reductive processes to bioremediation has been examined under denitrifying, sulfidogenic, and methanogenic conditions. An important observation

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was the dechlorination of TeCE to ethene in an enrichment culture when it was linked to the anaerobic degradation of toluene (Shen and Sewell 2005). Practical implementations have been discussed, and many references to conference proceedings and reports have been given (Lee et al. 1998). The following provide illustrations of effective procedures: 1. An enrichment culture KB-1 that contained organisms related to Dehalococcoides ethenogenes was used to treat groundwater contaminated with TeCE and lower concentrations of TCE (Major et al. 2002). The groundwater was first inoculated with methanol and acetate to produce anaerobic conditions before introduction of the culture. Within 200 d, the concentrations of all the chlorooethenes were low, and their loss was equivalent to the ethene that was produced. It was shown using a Dehalococcoides ethenogenes—specific PCR that the population increased during treatment and that this populated the contaminated aquifer. 2. A valuable comparison was made of the relative effectiveness of (a) augmentation with a culture containing Dehalococcoides ethenogenes, (b) stimulation of the natural dechlorinated population by addition of lactate and mineral nutrients, and (c) a recirculation control without amendment (Lendvay et al. 2003). The first procedure rapidly affected complete dechlorination, whereas the second was effective only after a lag period, and in both the first and second treatments the population of Dehalococcoides increased. 3. A comparison of electron donors was made to establish the effectiveness of the dechlorination of TCE to ethene. The results of analysis of 16S rDNA and terminal restriction fragment length polymorphism showed that five groups of organisms were found in all the enrichments including Dehalococcoides sp. whose presence was proved by the quantitative PCR. Although dechlorination of TCE to ethene has been demonstrated in species of this genus, other organisms including species of Citrobacter, Bacteroides, and Desulfovibrio probably played a role in establishing syntrophic associations (Freeborn et al. 2005). 4. The organisms in a reactor fed with TeCE and methanol were analyzed using a 16S rDNA clone library and fluorescence in situ hybridization (Rossetti et al. 2003). Dechlorination to chloroethene was the major reaction and, in contrast to the enrichments already noted and those that are discussed in Chapter 7, Part 3, sequences similar to those of the well-established dechlorinating organisms Dehalococcoides ethenogenes, Dehalobacter restrictus, Desulfuromonas chloroethenica, or Sulfurospirillum multivorans were absent. The majority of clones belonged to the phyla Spirochaetes (56%), Firmicutes (18%), Chloroflexi (16%), and Bacteroidetes (6%). No microbial type could therefore be correlated with the observed dechlorination, and it was concluded that dechlorination was carried out by hitherto undescribed organisms. Application of Higher Plants Use of higher plants has attracted attention for the remediation of TCE-contaminated aquatic systems, and in particular the use of hybrid poplar (Newman et al. 1999). Although this was encouraged by the removal of labeled TCE to >90% using hybrid poplar under hydroponic conditions (Orchard et al. 2000), and the ability of hybrid poplar to metabolize TCE (Newman et al. 1997; Shang et al. 2001), there are additional considerations that must be evaluated to determine its general applicability. a. The concentration of the established metabolites is generally low, and was apparently greatest in the roots (Orchard et al. 2000). This suggests the possible involvement of microbial activity. b. Loss by volatilization can occur from the stems by transport along the transpiration pathway (Ma and Burken 2003), and may present a problem through its introduction into the atmosphere.

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CONCLUSION In situ procedures using indigenous organisms induced for the aerobic degradation of chloroethenes have been successful, particularly in broadening the range of substrates to include those that might be produced by partial anaerobic dechlorination. Limited use has been made of anaerobic processes, though these may effectively be combined with aerobic reactions. The application of introduced organisms is currently limited by the relatively short time during which their activity can be maintained, but this could possibly be achieved by designing more effective “in situ bioreactors.”

METHYL TERT-BUTYL ETHER Methyl tert-butyl ether (MTBE) has replaced tetraethyl lead as an octane booster in gasoline and to minimize automobile discharge of carbon monoxide in cold weather. It is produced on a large scale, and there has been concern over its possible adverse effects. Its occurrence in groundwater can be associated with gasoline spillage or leakage from underground tanks that are used to store oxygenated fuel. This has already been discussed as a source for BTEX. A survey has revealed the widespread distribution of MTBE in water in the United States, although in no case did concentrations exceed those set for drinking water standards (Squillace et al. 1996). Contrary to initial fears, this compound was not associated with the simultaneous presence of BTEX so that its source remained unestablished. Attention has been directed to its toxicity, and a somewhat complex picture has emerged (references in Kado et al. 1998; Nihlén et al. 1998). Attention is drawn here only to some aspects of its biodegradability and the possibility of bioremediation of adversely affected groundwater. Experiments have been carried out under different conditions and reached apparently confl icting conclusions. 1. Under aerobic conditions, a mixed bacterial culture was able to degrade the substrate with intermediate formation of tert-butanol (Salanitro et al. 1994). 2. Evidence from field measurements indicated slow decay near the source of contamination but persistence further downstream. This was consistent with the results of laboratory experiments that showed only slow biodegradation of MTBE under aerobic condition (Borden et al. 1997). 3. In a simulated spill, groundwater was spiked with BTEX, MTBE, and chloride as marker, and injected into a sandy aquifer (Schirmer and Barker 1998). Whereas the level of BTEX fell, loss of MTBE was much slower, and only ca. 3% remained after 8 years, although the mechanism of its loss was not resolved. 4. Degradation of [U-14C]-MTBE and [U-14C]-tert-butanol was demonstrated within 105 d in laboratory micrososms containing stream sediments from a contaminated area (Bradley et al. 1994). 5. Under prolonged incubation under a variety of anaerobic conditions, the substrate was recalcitrant under sulfate-reducing conditions, and very poorly degraded under nitratereducing conditions (Mormille et al 1994). However, fission of the ether group to tertbutanol has been observed in laboratory sulfate-reducing cultures (Somsamak et al. 2001), and in methanogenic enrichment cultures (Somsamak et al. 2005), while extensive degradation in groundwater has been reported (Kuder et al. 2005). 6. As a result of the slow growth rates of organisms that can degrade MTBE, a membrane reactor was designed to retain the organisms while allowing free access of water within the system. In one application the inoculum was grown separately with MTBE and BTEX, and added as a concentrate to the reactor that was fed with a nutrient solution of ammonium sulfate and potassium phosphate. The influent contained MTBE, its degradation products (tert-butyl formate and tert-butanol), and BTEX. Over a period of 6 months the removal efficiency for the contaminants exceeded 99% (Zein et al. 2006).

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The collective evidence suggests the variable recalcitrance of MTBE under aerobic conditions, a degree of recalcitrance anaerobic conditions, and that the presence of the putative degradation product tert-butanol cannot necessarily be construed as evidence for biodegradation (Landmeyer et al. 1998). Bioremediation of groundwater contaminated with MTBE is clearly beset with problems for its practical implementation.

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Mars AE, GT Prins, P Wietzes, W de Konig, DB Janssen (1998) Effect of trichloroethylene on the competitive behaviour of toluene-degrading bacteria. Appl Environ Microbiol 64: 208–215. Mars AE, J Houwing, J Dfolfing, DB Janssen (1996) Degradation of toluene and trichloroethylene by Burkholderia cepacia G4 in growth-limited fed-batch culture. Appl Environ Microbiol 62: 886–891. Massol-Deyá A, R Weller, L Ríos-Hernández, J-Z Zhou, RF Hickey, JM Tiedje (1997) Succession and convergence of biofilm communities in fixed-film reactors treating aromatic hydrocarbons in groundwater. Appl Environ Microbiol 63: 270–276. McCarty PL, MN Goltz, GD Hopkins, ME Dolan, JP Allan, BT Kawakami, TJ Carrothers (1998) Full-scale evaluation of in situ cometabolic degradation of trichloroethylene in groundwater through toluene injection. Environ Sci Technol 32: 88–100. Morgan P, ST Lewis, RJ Watkinson (1993) Biodegradation of benzene, toluene, ethylbenzene and xylenes in gas-condensate-contaminated water. Environ Pollut 82: 181–190. Mormille MR, S Liu, JM Suflita (1994) Anaerobic biodegradation of gasoline oxygenates: extrapolation of information to multiple sites and redox conditions. Environ Sci Technol 28: 1727–1732. Munakata-Marr J, PL McCarty, MS Shields, M Reagin, CS Francesconi (1996) Enhancement of trichloroethylene degradation in aquifer microcosms bioaugmented with wild-type and genetically altered Burkholderia Pseudomonas cepacia G4 and PR1. Environ Sci Technol 30: 2045–2052. Nelson MJK, SO Montgomery, PH Pritchard (1988) Trichloroethylene metabolism by microorganisms that degrade aromatic compounds. Appl Environ Microbiol 54: 604–606. Neumann A, H Scholz-Muramatsu, G Diekert (1994) Tetrachloroethene metabolism of Dehalospirillum multivorans. Arch Microbiol 162: 295–301. Newman LA, SE Strand, N Choe, J Duffy, G Ekuan, M Ruiszai, BB Shurtleff, J Wilmoth, MP Gordon (1997) Uptake and biotransformation of trichloroethylene by hybrid poplars. Environ Sci Technol 31: 1062–1067. Newman LA, X Wang, IA Muizniekls, G Ekuan, M Ruszaj, R Cortelucci, D Domroes, G Karscig, T Newman, RS Crampton, MG Yost, PE Heilman, J Duffy, MP Gordon, SE Strand (1999) Remediation of trichloroethylene in an artificial aquifer with trees: a controlled field study. Environ Sci Technol 33: 2257–2265. Nihlén A, R Wålindrer, A Löf, G Johanson (1998) Experimental exposure to methyl tertiary-butyl ether. II. Acute effects in humans. Toxicol Appl Pharmacol 148: 281–287. Orchard BJ, WJ Doucette, JK Chard, B Bugbee (2000) Uptake of trichloroethylene by hybrid poplar trees grown hydroponically in flow-through plant growth chambers. Environ Toxicol Chem 19: 895–903. Rabus R, F Widdel (1995) Anaerobic degradation of ethylbenzene and other aromatic hydrocarbons by new denitrifying bacteria. Arch Microbiol 163: 96–103. Rabus R, F Widdel (1996) Utilization of alkylbenzenes during anaerobic growth of pure cultures of denitrifying bacteria on crude oil. Appl Environ Microbiol 62: 1238–1241. Rabus R, R Nordhaus, W Ludwig, F Widdel (1993) Complete oxidation of toluene under strictly anoxic conditions by a new sulfate-reducing bacterium. Appl Environ Microbiol 59: 1444–1451. Reinhard M, S Shang, PK Kitanidis, E Orwin, GD Hopkins, CA Lebron (1998) In situ BTEX biotransformation under enhanced nitrate- and sulfate-reducing conditions. Environ Sci Technol 31: 28–36. Rossetti S, LL Blackall, M Majone, P Hugenholtz, JJ Plumb, V Tandoi (2003) Kinetic and phylogenetic characterization of an anaerobic dechlorinating microbial community. Microbiology (UK) 149: 459–469. Safinowski M, C Griebler, RU Meckenstock (2006) Anaerobic cometabolic transformation of polycyclic and heterocyclic aromatic hydrocarbons: evidence from laboratory and field studies. Environ Sci Technol 40: 4165–4173. Sahoo D, JA Smith, TE Imbrigiotta, HM Mclallan (1998) Surfactant-enhanced remediation of a trichloroethene-contaminated aquifer. 2. Transport of TCE. Environ Sci Technol 32: 1686–1693. Salanitro JP (1993) The role of bioatttenuation in the management of aromatic hydrocarbon plumes in aquifers. Ground Water Monit Remed 13: 150–161. Salanitro JP, LA Diaz, MP Williams, HL Wisniewski (1994) Isolation of a bacterial culture that degrades methyl t-butyl ether. Appl Environ Microbiol 60: 2593–2596. Schirmer M, JF Barker (1998) A study of long-term MTBE attenuation in the Borden aquifer, Ontario, Canada. Ground Water Monitoring and Remediation: 113–122. Shang TO, SL Doty, AM Wilson, WNJ Howard, MP Gordon (2001) Trichloroethylene oxidative metabolism in plants: the trichloroethanol pathway. Phytochemistry 58: 1055–1065. Shen H, GW Sewell (2005) Reductive biotransformation of tetrachloroethene to ethene during anaerobic degradation of toluene: experimental evidence and kinetics. Environ Sci Technol 39: 9286–9294.

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Shih C-C, ME Davey, J Zhou, JM Tiedje, CS Criddle (1996) Effects of phenol feeding pattern on microbial community structure and cometabolism of trichloroethylene. Appl Environ Microbiol 62: 2953–2960. Somsamak P, HH Richnow, MM Häggblom (2005) Carbon isotope fractionation during anaerobic biotransformation of methyl tert-butyl ether and tert-amyl methyl ether. Environ Sci Technol 39: 103–109. Somsamak P, RM Cowan, MM Häggblom (2001) Anaerobic biotransformation of fuel oxygenates under sulfate-reducing conditions. FEMS Microbiol Ecol 37: 259–264. Squillace PI, JS Zogorski, WG Wilber, CV Price (1996) Preliminary assessment of the occurrence and possible sources of MTBE in groundwater in the United States, 1993–1994. Environ Sci Technol 30: 1721–1730. Stroo HF et al. (2003) Remediating chlorinated solvent source zones Environ Sci Technol 37: 225A–230A. Tick GR, F Lourenso, AL Wood, ML Brusseau (2003) Pilot-scale demonstration of cyclodextrin as a solubilityenhancement agent for remediation of a tetrachloroethene-contaminated aquifer Environ Sci Technol 37: 5829–5834. Townsend GT, JM Suflita (1996) Characterization of chloroethylene dehalogenation by cell extracts of Desulfomonile tiedjei and its relationship to chlorobenzoate dehalogenation. Appl Environ Microbiol 62: 2850–2853. Tsao C-W, H-G Song, R Bartha (1998) Metabolism of benzene, toluene, and xylene hydrocarbons in soil. Appl Environ Microbiol 64: 4924–4929. Vanderberg LA, BL Burback, JJ Perry (1995) Biodegradation of trichloroethylene by Mycobacterium vaccae. Can J Microbiol 41: 298–301. Vannelli T, M Logan, DM Arciero, AN Hooper (1990) Degradation of halogenated aliphatic compounds by the ammonia-oxidizing bacterium Nitrosomonas europaea. Appl Environ Microbiol 56: 1169–1171. Walter GA, SE Strand, RP Herwig, TP Treat, HD Stensel (1997) Trichloroethylene and methane feeding strategies to sustain degradation by methanotrophic enrichments. Water Environ Res 69: 1066–1074. Zein MM, MT Suidan, AD Venosa (2006) Bioremediation of groundwater contaminated with gasoline hydrocarbons and oxygenates using a membrane-based reactor. Environ Sci Technol 40: 1997–2003.

FINALE: A HIERARCHICAL STRATEGY It is obvious that there are several basic questions that must be answered before considering bioremediation: • • • • •

What compounds are present and which have to be removed? What quantities are involved? By what means is success to be evaluated? Are indigenous or exogenous organisms to be used? What additional carbon additives are appropriate?

In virtually all the examples that have been used as illustration, three stages may be discerned, and it is convenient to assemble these into a hierarchical system: (I) basic laboratory studies, (II) microcosm studies using site material, and (III) evaluation in large-scale systems, on-site pilot plants, or in situ installations. For all of them, the use of specific isotope enrichment (Chapter 13) can provide invaluable evidence of degradation or transformation. It is essential in addition that geological, engineering, microbiological, and chemical interactions be setup before initiating the project and maintained throughout. The design of laboratory experiments on degradation has been discussed in Chapter 5 and includes illustrative examples of both micro- and mesocosm systems. Several examples of the use of microcosms have been given in this chapter.

I. LABORATORY EXPERIMENTS For both exogenous and indigenous organisms, the following merit attention: • Examination of pathways including transformation with the formation of intermediate and terminal metabolites

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• • • • •

Use of the substrates as sources of carbon, nitrogen, sulfur, or phosphorus Physiological optima of temperature, pH, and oxygen concentration Toxicity of the substrates Induction of catabolic enzymes by growth with structurally unrelated substrates Degradation of a compound during growth with a metabolically compatible substrate (cometabolism or concurrent metabolism) • The existence of mixtures of substrates when their biodegradation can be metabolically incompatible • Evaluation of relevant probes for the presence of the relevant microorganisms and their activity For many exogenous organisms, many of these issues will already have been established, so that only complementation will be required for indigenous organisms: determination of their presence, number, and metabolic activity by conventional laboratory procedures. It may be valuable to set up enrichment cultures using site material to provide suitable material for laboratory and microcosm studies.

II. MICROCOSM EXPERIMENTS These should be constructed from material collected from the relevant site and simulate aerobic, microaerophilic, or anaerobic conditions. They should provide the kinetic data necessary for largescale operation, and specifically address the following—although some of them will already have been examined: • Rates of substrate loss • Data on the existence and stability of metabolites and their toxicity to other biota • Existence of threshold substrate concentration below which rates of degradation are low or negligible • Effectiveness of analog substrates and metabolites in promoting degradation • Stability of the system under prolonged operation It is appropriate at this stage to evaluate procedures that will be used to assess the effectiveness of bioremediation, which have been discussed in Chapter 13. These may include (a) use of radiolabeled substrates (although these will not generally be permitted in field operations) and the application of 13 C-labeled substrates, (b) evaluation of the occurrence of metabolites, and (c) evaluation of markers such as specific enzymes.

III. LARGE-SCALE OPERATIONS Apart from the preceding issues, it is specially important at this stage that very close collaboration with geologists and engineers is necessary. Putatively optimal solutions may not be realistic for the site, and alternatives may have to be adopted. The choice may then depend on returning to Stage II—or even Stage I—to evaluate these new constraints. A good example is provided by the reduced permeability of substrata encountered during application of a pump-and-treat system: this was due to several causes including mobilization of colloidal material in the system and subsequent deposition and fouling at injection sites and infiltration zones (Wiener et al. 1996). It is probably inadvisable to initiate the evaluation of alternative strategies in large-scale systems. All the intrinsic problems of scale-up will clearly have to be addressed and solved. These include: (a) the mobility of microorganisms, substrates, nutrients, and metabolites, (b) oxygen transport or the maintenance of other electron acceptors generally nitrate, and (c) the long-term stability of the system.

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CONCLUDING COMMENTS Whereas the potential of bioremediation is substantial, its application may encounter important limitations that are apparent from many of the examples that have been used as illustration. The authors feel that these limitations can be overcome only when adequate attention is directed to fundamental microbiological, chemical, and engineering issues. Only then can the barriers to achieving success in full-scale operation be surmounted, since there are significant problems with in situ or on-site programs that have not taken advantage of experiments under the controlled conditions in microcosms or small-scale bioreactors. Of the several unresolved microbiological issues, it is suggested that the following deserve particular consideration: • The degree to which the compounds are accessible to the appropriate microorganisms. • The extent to which the population of degradative organisms can be maintained and increased. • The relative biodegradability of components in complex mixtures such as PAHs and PCBs, since some of the less readily degraded components may be the least desirable from an environmental and toxicological viewpoint. • Even when removal of, say 90% of the contaminant has been achieved, it may not be possible to attain 99% removal due to the possible existence of threshold concentrations below which rates of degradation are slow or even negligible. • Evaluation of success is of primary importance. Loss of substrates is a necessary but not sufficient condition, in the light of the frequency of biotransformation reactions and the formation of terminal and possibly toxic metabolites.

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Index A Acenaphthene-[13C] aerobic degradation, NMR application, 286 Acetaminophen, chlorination (Cl2), 33 Acetate anaerobic metabolism/dismutation, 319 methanogenesis, 319 Acetone, aerobic metabolism, 306 Acetophenone, degradation, 434 Acetylene hydratase, 187, 309 Acid tolerance, amino acid decarboxylation, 176 Actinomycetes, 65 cytochrome P450 monooxygenation, 65 dioxygenation, 66 Acyl chloride intermediates, 122, 126, 455–456 chlorobenzene degradation, 122 hexachlorocyclohexane degradation, 122 pentachlorophenol degradation, 122 Adamantane, cytochrome P450 hydroxylation, 115 Adaptive degradation, 216−217 Adenine, aerobic degradation, 542 Aging determinant biodegradation, 208–210 bioremediation, 611 Agrochemicals, adaptive degradation, 217 Alachlor, chironomid larvae transformation, 96 Aldehyde oxidoreductase, 164 Aldicarb, chemical hydrolysis, 23 Aliphatic ethers, aerobic degradation, 573 Alkaliphiles, 75 Alkanes aerobic degradation, 298−303, 299 chain-branched, 300, 301 cytochrome P450 bacteria, 303 Prototheca zopfii, 99 yeast, 132, 303 anaerobic degradation, 304, 305 atmospheric transformation, 17, 18 bioremediation, 639−642 bioavailability, 640 surfactant production, 640 terrestrial habitat requirements, 640 oxygen provision, 640 inorganic nitrogen source, 640 growth, emulsifier, Candida, 213 oxidation, Nitrosomonas europaea, 71 Alkane degradation, marine systems, 641 nutrient addition, 642 oil dispersant, 642 Alkane hydroxylase, 103, 301, 303 ω-alkane hydroxylase, 300 Alkane hydroxylation, 298−299, 299 yeast, 116 Alkane monooxygenase gene, rhizosphere, 652 Alkane sulfonates, electron acceptor, 156 Alkanones, degradation, 310

Alkenes double bond fission, 305 epoxidation, 104, 305−308 Alkylethoxy sulfates degradation, ether bond fission, 570 Alkyl arenes, atmospheric transformation, 17 Alkyl carboxylates aerobic reduction, 164 Alkyl phosphonates, reduction alkanes, 589 Alkyl sulfatases, 569 Alkyl sulfates, hydrolysis/inversion alkyl−oxygen fission, 570 Alkyl sulfides, atmospheric transformation, 21 Alkyl sulfonate, anaerobic degradation, Bilophilia, 590 Alkylsulfonate-[35S], metabolism, 279 Alkylamines deaminase, 313 degradation, oxidase, 132 dehydrogenase, 311 oxidase, 311 Alkylbenzenes, degradation, 149 Alkylnaphthalenes, aerobic bacterial oxidation, 400 Alkylphenol polyethoxylates, partial degradation, 575 Alkynes, triple bond hydration, 308 Alternative substrates, growth, 258 Amidase, 322 Amino acid metabolism anaerobic rearrangement, 316 dismutation (Stickland reaction), 316 decarboxylation, 315 fermentation, clostridia, 73 pyridoxal 5′-phosphate enzymes, 312 pyruvoyl enzymes, 315 2-aminobenzenesulfonate, degradation, 507 4-aminobenzenesulfonate degradation, 507 excretion, 521 Aminobenzoate monooxygenation, 109 2-aminobenzoate, anaerobic degradation, 436 2-aminobenzoate CoA ester, aerobic degradation Azoarcus, 432, 433 3-aminobenzoate-6-hydroxylase, 109 1-aminocyclopropane-1-carboxylate (ACC) deaminase, 609 degradation, 313, 314 Aminodinitrotoluene, ozone degradation, 30 2-aminodiphenylamine, transformation phenazine, 4-aminoacridine, 510 2-aminoethylarsonic acid, degradation, 594 2-aminoethylphosphonate, degradation, 588 2-aminoethylsulfonate, dioxygenation, 126 5-aminonaphthalene-2-sulfonate, sequential microbial/ chemical reactions, 55 6-aminonaphthalene-2-sulfonate, degradation, 507 2-aminophenol distal 1,6-dioxygenase, 124, 511 2-aminophenoxazine-3-one, terminal metabolite, 510 2-aminopropane, transformation to alaninol, 311 5-aminosalicylate degradation, 426

693

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694 Amitrole (3-amino-1H-1,2,4-triazole), TiO2 photochemical degradation, 10 Anaerobic bacteria carboxylation, 449−452 experimental procedures, 258 nitrate tolerance, 72 oxygen tolerance, 72, 202 Anaerobic xylene degradation, methylbenzylsuccinate markers, 681 Anaeromyxobacter, 2-chlorophenol anaerobic dechlorination, 488 Anammox, 148 Androgens, cytochrome P450 demethylation, 117 Aniline aerobic degradation, 452 dioxygenase, 452 ozone transformation, 31 Aniline-[15N] reaction humic/fulvic acid, NMR application, 287 ANME-1/ANME-2, methane anaerobic oxidation, 626 Anodic oxidation, 37 Anthracene degradation bacteria 403, 404 white-rot fungi, 414 Anthracene-9[13C] degradation, 278 Anthranilate 1,2-dioxygenase, 425 Anthranilate-5-hydroxylase, 108 Anthraquinone-2,6-disulfonate (AQDS) as electron acceptor, 153−155 Antibiotic resistance, 170−172 covalent modification, 170 efflux pump, 171 plasmids, 224 Aquatic plant metabolism pentachlorophenol, 604 phoxim, 604 2,4,6-trinitrotoluene (TNT), 604 Archaea, 58 Archaeal population, oil contamination, 623 Arene aldehyde oxidoreductase, Clostridium, 439 Arene aldehyde, anaerobic reactions oxidation/sulfate reduction 439 reduction to benzyl alcohols, 439 Arene carboxylate degradation, low oxygen concentration, 204 Arene carboxylates. See also specific compounds aerobic reduction to aldehyde, 433 decarboxylation, 430 low oxygen concentration, degradation, 203 Arene cis-dihydrodiols, synthetic applications, 392−394, 393 Arene dihydroxylation, 18O application, 279 Arene hydrocarbon contamination, 13C-naphthalene analysis, 627 Arene hydrocarbons. See also PAH degradation, Novosphingobium, 67 δ13C fractionation, 630 pyrolysis, 35 Arene hydroxylation fungi, 389 Streptomyces, 115 cytochrome P450, 65, 388 Arene monooxygenation, 106

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Index Arene oxide, NIH shift, 107, 410, 411 Arenesulfonates anaerobic desulfonation, 508 dioxygenase, 506 Arginine, anaerobic growth Pseudomonas, 201 Aromatic amino acids, sequential microbial/chemical reactions, 55 Aromatic carboxylic acids, decarboxylation, 68 Arsenate, electron acceptor, 151 Arsenoacetate, growth substrate, 594 Artificial streams, pentachlorophenol degradation, 266 Aryl carboxylates, aerobic reduction, 164 Aryl sulfatase, mycobacteria, 570 Aryl sulfates, hydrolysis with O−S fission, 570 Aryl-alcohol oxidase, fungi, 133 Arylamine reactions N-acetylation, 509 association with organic matter, 509 dimerization, 509 Association organic compounds humic matter, 15 N − NMR application, 287 Association, substrates − soil organics, 206 Atmospheric reactants − hydroxyl radicals, nitrate radicals, ozone, 14 Atmospheric reactions relative rates, 15 survey, 15 ATP synthesis anaerobic bacteria, 52 purine degradation, 544 coupled to transformation 3-chlorobenzoate, 475 hexachlorocyclohexane, 355, 369 tri-/tetrachloroethene, 367 Atrazine bioremediation, carbon/nitrogen ratio, 673 degradation, 545, 546 metabolism, hybrid poplar, 98, 603 ozone/H2O2 transformation, 31 photochemical transformation, 5, 9 sulfate radical degradation, 32 Atropine, degradation, 345, 346 Azaarene. See also specific compounds inhibition PAH degradation, 647 Azoarene reductase, 163−164 aerobic bacteria, 521 Clostridium perfringens, 521 reduction/N−N scission, 520 Azo group decolorization fungal peroxidase, 521 yeast, 521 Azo pigments/dyes, reduction/N−N scission, 507 Azotobacter sp., 2,4,6-trichlorophenol degradation, 66 Azotobacter, degradation potential, 66

B Bacterial dioxygenases, 105 Baeyer-Villiger monooxygenation acetophenone, 434 alkanone, 310 cycloalkanone, 337

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Index Bentazone, association soil organic matter, 207 Benz[a]anthracene, bacterial metabolism, 406 Benzene aerobic degradation, 386 anaerobic degradation, δ13C/δ2H, 630 anaerobic hydroxylation, 18O application, 280 cytochrome P450 hydroxylation, 115 dioxygenase, 387 metabolism, spinach, 604 oxidation, Nitrosomonas europaea, 71 monooxygenation, 387 Benzene-[13C] degradation, 278 Benzene-[13C]/toluene-[13C], δ13C lipid analysis, 630−631 Benzoate/acetate, diauxic growth, 221 Benzoate. See also specific compounds anaerobic degradation, 436−438, 437 dioxygenase, 424 metabolism, Trichosporon, 427 transport, 214 Benzofuran dioxygenation, 555 hydroxylation/ring fission, Paecilomyces, 558 Benzo[a]pyrene metabolism bacteria, 408 Cunninghamella transformation, 412 cytochrome P450 Saccharomyces hydroxylase, 413 Paralichthys, 93, 94 photooxidation, 11 quinones, white-rot fungal metabolites, 416 Saccharomyces hydroxylation, 116 Selenastrum, 413 transformation, white-rot fungi, 416 Benzo[a]pyrene quinones, white-rot fungal metabolites, 416 Benzo[b]thiophene, sequential microbial/chemical reactions, 55 Benzothiazole degradation, 567, 568 [15N] degradation, NMR application, 287 Benzothiophene sulfoxide, abiotic reaction, 566, 567 Benzothiophene dioxygenation, 565 oxidation sulfoxide, 566 Benzoyl CoA ester aerobic degradation, Azoarcus, 432 pathway inducer, 436 reductase, Thauera aromatica, 290 Benzylsuccinate synthase, 450 toluene, anaerobic metabolism, 390−391 Bias in elective enrichment, 622 Bile acids, anaerobic transformation, 343 Biocontrol agents, bacteria, 607 Biodegradability, ready/inherent, 247 Biodegradation/biotransformation, definitions, 51 Biofilms bioremediation, 611 BTEX degradation, 680 Biogenic terpenes, atmospheric transformation, 17 Bioremediation, pathway regulation, 610 Biosurfactant production alkane growth, 640 PAH growth, 649

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695 Biphenyl aerobic degradation, 401 Porphyrobacter, 59 dioxygenase, 401 hydroxylation/ring fission, Paecilomyces liliacus, 77, 402, 558 inducer, PCB degradation, 196, 210 regulation degradation pathway, 464 transformation, Cunninghamella, 411 Bis-(1-chloro-2-propyl)ether, enantiomeric biodegradation, 54, 575 Bis-(pentafluorophenylpropyl)sulfide, sulfur source, 578 Bisphenol-A aerobic degradation, 301, 302, 447 cytochrome P450 hydroxylation, 116 Boronate, monooxygenation, 591 Box proteins, aerobic benzoyl-CoA degradation, 432 Bradyrhizobium, constitutive 3-ketoadipate enzymes, 221 Brominated biphenyls, primers anaerobic PCB dechlorination, 196, 666 2-bromobenzoate, aerobic degradation, 469−470, 470 4-bromobenzoate, primer anaerobic PCB dechlorination, 467, 666 5-bromo-5,6-dihydrouracil, debromination, 541 2-bromodibenzo[1,4]dioxin, primer octachlorodibenzo [1,4]dioxin anaerobic dechlorination, 467, 666 5-bromouridylate, debromination, 541 Bromoacetaldehyde toxicity, 360 Bromophenol, reductive debromination, Desulfovibrio, 488 BTEX aerobic degradation, 679−680 aerobic metabolites, humic associated, 680 13 C-benzene analysis, 628 fluidized-bed reactor, 680 white-rot fungus, 389 BTEX, groundwater contamination, 679−681 anaerobic conditions, 680−681 relative component degradation, 681 methylbenzylsuccinate markers, xylenes, 681 in situ stream treatment, 680 remediation advantage, indigenous organisms, 679 toluene/biphenyl dioxygenase family, 623 tert-Butyl diphenyl phosphate, degradation, 265

C 13

C Metabolic application, 277−278 NMR, 285−286 C14 Methyl steroid oxidation, 18O application, 280 Camphor catabolic plasmid, 225 cytochrome P450 (P450cam) hydroxylation, 115, 458 degradation, 340 Carbamazepine, ozone degradation, 30 Carbaryl, abiotic transformation, 3 Carbazole 1,9a dioxygenase, 555 three component, 527 relaxed specificity, 527 Carbazole, aerobic degradation, 527, 528 Carbocyclic ring reductase, 164 Carbon disulfide, growth substrate, 580 Carbon monoxide dehydrogenase, 183, 185 structure, 291

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696 Carbon monoxide metabolite, 128, 539, 559 Carbon monoxide metabolism, 323 Carboxylation alkane degradation aerobic, 300 anaerobic, 304 phenol/aniline anaerobic, 449−452 Carboxylic acid reductase, 188 Carotenoids, sulfide reduction, 28 Caryophyllene oxide, transformation, Botrytis, 345 Catabolic plasmids, 225−226, 225 Catechol alternative ring fission, 123 degradation, 472−474 incompatibilities, 3-substituted, 474 ring fission dioxygenation, 428. See also Extradiol ring fission, Intradiol ring fission Catechol 2,3-dioxygenase gene, rhizosphere, 652 Cation electron aceptor, 152−154 reduction, 172 toxicity efflux system, 175 polyphosphate, 175 Chemical industry contaminants, conclusion for bioremediation, 668 Chlorate electron acceptor, 150 reductase, 150−151, 187 Chlorinated alkane degradation, Rhodococcus, 64 Chlorinated arene hydrocarbons. See also specific compounds atmospheric transformation, 17 remediation, δ13C application, 631 Chlorinated biphenyl, catabolic plasmid, 225 Chlorinated dibenzo[1,4]dioxin partial anaerobic dechlorination, 558 photoproduction, pentachlorophenol, 7 thermal production, 35 Chlorinated ethanes, zero-valent iron transformation, 26 Chlorinated phenoxyalkanoates, remediation, 672. See also specific compounds Chlorination (Cl2 ), 32 Chlorite dismutase, 151 Chloroacetophenone, PCB metabolite, 463 2-chloroacetophenone degradation, 18O application, 280 2-chloroacrylate degradation, 362 reductase, 163 3-chloroacrylate, degradation, 363 Chloroalkane/chloroalkene, δ13C, δ37Cl, 632 Chloroalkanes oxidation, Nitrosomonas europaea, 71, 364 Chloroalkene, aerobic degradation, δ13C application, 631 Chloroaniline, regulation of bacterial degradation, 672 4-chloroaniline aerobic degradation, 490 association with soil organic matter, 207 Chloroanisole metabolite, 659 Chlorobenzene/toluene degradation Phanerochaete, 610 Ralstonia, 662−663 Chlorobenzene degradation. See also specific compounds

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Index acyl chloride intermediate, 122 Fenton’s reagent transformation, 9 terminal metabolite hexachlorocyclohexane, 662 threshold concentration, 212 Chlorobenzoates marker, PCB metabolism, 664 PCB metabolites, 463 3-chlorobenzoate aerobic degradation, 469, 470−472 low oxygen concentration, 203 pristine environment, 216 anaerobic dechlorination, Desulfomonile, 475 catabolic plasmid, persistence, 227 3-chlorobenzoate-degrading genotypes, 625 4-chlorobenzoate aerobic degradation, 469, 470−472 transport, 214 2-chlorobiphenyl metabolism, NIH shift, 464 4-chlorobiphenyl, nitrated metabolites, 57 3-chlorocatechol aerobic metabolism ring fission, 469, 471 alternatives, 123, 456 Candida, 487 toxic metabolite, 222 3-chlorocatechol 1,6-dioxygenation 470, 471 Chlorocatechol-sediment, resistance to anaerobic dechlorination, 209 3-chloro-/3-methylcatechol, degradation incompatibility, 223 4-chlorocatechol, aerobic metabolism, 469, 471 3-chloro-1,2-epoxyethane (epichlorohydrin), degradation, 362 Chloroethene anaerobic dechlorination, 684−685 cooxidation, 682 reductase, 368−369 Chloroethene (vinyl chloride) aerobic degradation, enrichment factors, 631 Chloroethene bioremediation in situ procedure preference, 682 application of exogenous bacteria, 684 application of indigenous bacteria, 683−684 toluene/phenol oxygenase induction, 682, 683 Chloroethene oxidation methane/propane/ammonia monooxygenase, 682 pseudomonad monooxygenase/dioxygenase, 682–683 Chlorofluoroalkanes, porphyrin dehalogenation, 27 Chloroform degradation, 364 methane/butane induction, 197 2-chlorohydroquinone dioxygenation, 126 5-chloro-2-hydroxypyridine-3-carboxylate, aerobic degradation, 534, 535 Chloromuconate metabolism, 470, 471 Chlorophenols. See also 2,4-Dichlorophenol, Pentachlorophenol aerobic bacteria, degradation, 482−483 anaerobic dechlorination, 488, 660 aerobic degradation mycobacteria, 63 Phanerochaete, 486−487 Sphingobium, 67 white-rot fungi, 659 metabolism, Candida, 487

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Index monooxygation, 111 reductive dechlorination, 488−489 remediation, 659 tolerance, 176 2-chlorophenol, electron acceptor, Anaeromyxobacter dehalogenans, 74 4-chlorophenol/4-methylphenol degradation, Comamonas, 610 Chlorophenoxyacetate, catabolic plasmid, 225 4-chlorophenylacetate, two-component dioxygenase, 475 2-chloropropionate dehalogenase, 362 5-chloropyrogallol, ring fission Candida, 487 3-chloroquinoline-8-carboxylate, dioxygenation, 537 Chlorotoluenes, transformation, 457 2-chlorotoluene, degradation, 456 Chlorpyrifos chlorination (Cl 2 ), 33 transformation, Enterobacter, 69 2-chloro-6-(trichloromethyl)-pyridine metabolism, Nitrosomonas europaea, 71 oxygen dependent metabolism, 202, 205 transformations, 534, 535 Cholesterol degradation, 149 oxidase oxidation, 132 Cholic acid biotransformation / biodegradation, oxygen dependence, 205 Chromium (VI) reduction, Deinococcus, 154 Chromous chloride dehalogenation, 26 Ciprofloxacin, fungal transformation, 539 Cis-stilbene, photocyclization, 11 Clinical bacteria, degradative capability, 58. See also Mycobacteria Clostridia, 73 CO2 alkanone degradation, 310 epoxide metabolism, 306 growth requirement, 53 Coal-tar globules, 646 Coenzyme M, vinyl chloride degradation, 307 Combined biological/chemical treatment, PAHs, 645 Cometabolism, 195 Complexing agents, degradation, 312 Continuous culture, application, 262 Cooxidation, chloroethene remediation, 682 Copper dioxygenase, 124 structure, Aspergillus, 290 Copper enzymes, 184−185 Copper methane monooxygenase, 184, 298 Copper−quinoprotein amine oxidase, structure, 291 C-P lyase, 588 Cresol degradation aerobic, 446 anaerobic, 451 fumarate pathway, 450 Cyanide, metabolism, 323 Cyanobacteria hexachlorocyclohexane dechlorination, 61 PAH metabolism, 413 Cycloalkanes, monooxygenation, 327 Cycloalkanone monooxygenase, 105, 106, 337

CRC_7241_Index.indd 697

697 Cycloclasticus, arene degradation, 58−59 Cyclodextrin, chloroethene solubility enhancement, 682 Cyclohexane aerobic degradation, 337 monooxygenase, 69, 70 Cyclohexenone, dismutation, 163 Cyclohexylacetate, degradation, 337, 338 Cycloisomerase limitation, degradation of 2-chlorotoluene, 456 Cyclopropane carboxylate, degradation, 337, 338 Rhodococcus, 65 Cytidine, aerobic degradation, 540 Cytochrome oxidase, anaerobic bacteria, 202 Cytochrome P450 monooxygenase fish, 92 oxaarene transformation, 553, 554 systems, 114 yeast, 116 oxygen level dependence, 204 Cytochrome P450cam, 115, 340, 458 l-cysteate, transformation, 590 l-cysteine, desulfhydrase, 579

D Daidzein, anaerobic degradation, 560, 562 DDE dioxygenation, 458 recalcitrance, 198 DDT CaO degradation, 29 degradation, Phanerochaete chrysosporium, 76 reduction, zero-valent iron, 25 resistance, flies, 96 ring hydroxylation, 458 transformation, 67, 350, 351 aquatic plants, 663 photochemical, 8 uptake/metabolism, plants, 663 Decabromodiphenyl ether metabolism anaerobic debromination, 468, 666 fish, 95 n-decane, anaerobic degradation, 304 Decarboxylation non-oxidative, 430−431 phthalate degradation, 431 reversible anaerobic, 431−432 Degradative population, molecular procedures, 622 Dehalogenases halogenated alkanes, 359−361 halogenated alkanoates, 362 reductive 3-chlorobenzoate, 475 tetrachloro-/trichloroethene, 367, 368, 684 Dehalorespiration, 156, 367 Dehydroabietic acid cytochrome P450 hydroxylation, 115 degradation, 344 Dehydroxylation, phenol anaerobic degradation, 165, 187, 449 Denitrification, 148 Dense cell-suspensions, 261 O-demethylase, three-component anaerobic, 577

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698 O-demethylation methoxybenzoates, 577 sulfide-dependent, 578 Desulfitobacterium spp., chlorophenol anaerobic dechlorination, 488 Desulfomonile, anaerobic dechlorination 3-chlorobenzoate, 475 chlorophenols, 488 Desulfovibrio vulgaris, oxygen tolerance, 202 Desulfurization, dibenzothiophene, 565 3,5-diacetamino-2,4,6-triiodobenzoate, partial deiodination, Trametes, 476 2,4-diacetylphloroglucinol hydrolase, 449 2,4-diacetylphloroglucinol, bacterial metabolite, 607, 608 α,ω-diamines, degradation, 311 2,3-diaminonaphthalene, reaction with nitrite, Escherichia coli, 55 Dibenzo[1,4]dioxan, degradation, 556, 557 Dibenzofuran degradation, 556, 557 Porphyrobacter, 59 dioxygenation, 556 hydroxylation/ring fission, Paecilomyces, 558 Dibenzothiophene anaerobic transformation, 567 dioxygenation, 565 monooxygenation, 113 nitrated metabolites from nitrite, 57 sulfur elimination, 565, 566 2,6-dibromobiphenyl, primer anaerobic PCB dechlorination, 465 Dibromoethane, methanogen anaerobic elimination, 355 Dibromoethene, methanogen anaerobic elimination, 355 4,4′-dicarboxyazobenzene, long-term adaptive degradation, 217 α,ω-Dicarboxylic acids, 300 3,4-Dichloroaniline association with soil organic matter, 207, 208 degradation bacteria, 490, 672 fungi, 673 nitrite reactions, Escherichia coli, 55 1,4-dichlorobenzene, degradation, 457 2,4-dichlorobenzoate, degradation, 472 reductive/hydrolytic steps, 475 2,7-dichlorodibenzo[1,4]dioxin degradation peroxidase, 559 Phanerochaete chrysosporium, 76 2,3-dichlorodibenzofuran, dioxygenation, 558 2,8-dichlorodibenzofuran, dioxygenation, 558 Dichlorodifluoromethane, CaO decomposition, 30 1,2-dichloroethane, aerobic dechlorination, 37Cl kinetic isotope effect, 633 1,1-sichloroethene, adverse effect on trichloroethene remediation, 684 2,6-dichlorohydroquinone degradation, 484 dioxygenation, 483 Dichloromethane aerobic degradation, fractionation factors 13 C / 37Cl, 632 degradation, 358, 363

CRC_7241_Index.indd 698

Index 2,4-dichlorophenol association with soil organic matter, 207, 208 degradation, 112, 483 Phanerochaete, 487 hydroxylase, 111 sulfate radical, degradation, 32 2,4-dichlorophenoxyacetate degradation plasmid/chromosomal genes, 226 pristine environment, 215 protection by sorption, 208, 210 threshold concentration, 212 dioxygenation, 127 Fenton’s reagent transformation, 9 metabolism, Aspergillus, NIH shift, 486, 487 transport, 214, 215 Dichlorophenoxypropionate degradation, Sphingobium, 67 1,3-dichloropropene, degradation, 362 2,2-dichloropropionate degradation, 362 dehalogenase bias, 626 Diclofenac, photochemical degradation, 8 Diethyl ether fungal degradation, Graphium, 575 monooxygenation, 574 Difluorobenzoate dioxygenation, 498−499, 499 19 F − NMR application, 288 2,2′-difluorobiphenyl, dioxygenation, 495 4,4′-difluorobiphenyl ether, dioxygenase transformation, 495 2,3-difluorophenol, transformation, Rhodococcus, 500, 501 Dihydropyrimidine dehydrogenase, dehalogenation, 541 Dihydroxybenzoates, anaerobic degradation, 438 3,4-dihydroxybenzoate aerobic degradation, regulation, 220 2,3-dihydroxybiphenyl 1,2-dioxygenase, 461 1,2-dihydroxynaphthalene dioxygenase, 507 2,5-dihydroxypyridine, degradation, 530, 531 7,12-dimethylbenz[a]anthracene, photooxidation, 11 2,2-dimethylpropionate, aerobic degradation, 301 Dimethyldisulfide, degradation, 580 Dimethylsulfide, anaerobic growth, 578 Dimethylsulfoniopropionate anaerobic degradation, 579−580 degradation, roseobacters, 59 degradation/elimination, 579 de-S-methylation, 580 Dinitramine dealkylation by carp, 93 fungal metabolism, 509 4,6-dinitrohexanoate, 2,4-dinitrophenol metabolite, 516 2,4-dinitrophenol aerobic degradation, 164 transformation, 516−517, 517 2,4-dinitrotoluene degradation 18 O application, 279 Phanerochaete chrysosporium, 513, 514 2,6-dinitrotoluene, nitro group reduction by Salmonella, 68, 509

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Index 2,4-/2,6-dinitrotoluene degradation, 676 dioxygenation, 512 Dinoseb, soil slurry remediation, starch waste/ pilot-scale inoculum, 673−674 1,4-dioxane, aerobic degradation, 576 Dioxygenases,104, 105 arene dihydroxylation, 118−122 concomitant elimination, 119–120, 120 2-halobenzoate 1,2-dioxygenases, 425, 496 two component, 469 three component, 470 2-ketoglutarate, 125−126 nitroarene three component, 512 PCB degradation, three component, 459 region III and IV in large α-subunit, 460 increased specificity, 460 Phanerochaete, 1,2,4-trihydroxybenzene, 129 phenol formation, 119 phenoxyalkanoate degradation, 577 ring fission, 123−125, 455−456 Dioxygenation pathways, PAH degradation, 399, 402, 404 Diphenyl ether degradation, 555−556, 555 Trichosporium, 78 Dismutations acetate, anaerobic metabolism, 319 amino acids (Stickland reaction), 317 cyclohexenone, 163 2-hydroxypurine, 544 trichloroacetaldehyde, 364 Dissolved organic carbon, chemical hydrolysis rate, 25 DMSO reductase family, 148, 187

E EDTA anodic oxidation, 30 degradation, 312 photochemical transformation, 7 transport, 215 Efflux systems antibiotic resistance, 171 hydrocarbon tolerance, 169 metal/metalloid resistance, 175 Elective enrichment, 249−252 Electrohydrolic discharge plasmas, 37 Electron acceptors alkane sulfonate, 156 arsenate, 151 chlorate, 150 2-chlorophenol, 74 humic acid, 155 metal cations, 152−154 nitrate, 148 nitroalkane, 156 selenate, 151 tetrachloroethene, 74 Electron acceptors, fractionation factor in toluene anaerobic degradation, 630 Electron donors, anaerobic trichloroethene dechlorination, 685 Elimination reactions, 350

CRC_7241_Index.indd 699

699 Elongation of carbon chain during alkethoxy sulfate degradation, 570 Emulsifier/biosurfactant, 213 Enantiomeric biodegradations bis-(1-chloro-2-propyl)ether, 54, 575 α-hexachlorocyclohexane, 54, 353 mandelate, 54, 435 phenoxyalkanoate, 577 3-phenylbutyrate, 54 2-(2-methyl-4-chlorophenoxy)propionate (Mecoprop), 54 β-endosulfan, transformation, 570 Enediol dioxygenase, 128, 182, 183 dual Fe and Ni enzymatic activity, 182 Enoate, anaerobic reduction, 163 α,β-Enone dismutation, 339 reduction clostridia, 73 flavanoid degradation, 560 reductase, clostridia, 163 Mössbauer spectrum, 280 Enrichments, elective bias, 622 Enrofloxacin, fungal transformation, 539 Enterobacteriaceae metabolism, 67 Enzyme-mediated substrate association, 206−208, 207 Enzymes, plant introduction 3-chlorocatechol 1,2-dioxygenase, 606 cytochrome P450 2E1, 606 extradiol catechol dioxygenase, 606 haloalkane dehalogenase, 606 Epichlorohydrin (3-chloro-1,2-epoxyethane), degradation, 362 Epoxidation, alkene, 305 Epoxides metabolism, 306 resolution, 307−308 EPR applications, 289 Estrogens, ozone degradation, 31 Estrone, chlorination (Cl2), 34 Ethionamide (pyridine-2-ethyl-4-carbothiomaide), metabolic activation, 170 Ethylbenzene aerobic hydroxylation, 388 anaerobic degradation, 391 4-ethylcatechol, toxic metabolite, 222, 224 4-ethylphenol dehydrogenation, 447 hydroxylation, Nitrosomonas, 388 EUROPHORE reactor, 21, 246 Extracellular enzymes, 213 peroxidase, white-rot fungi, 133 Extradiol ring fission, 105, 110, 123 dioxygenases, 123 Fe3+ enzyme, 428 Mn enzyme, 428

F F420 dependent reductase, 164 hydride transfer, ring reduction, 516 Fe(III) reductase, Geobacter sulfurreducens, 153, 155

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700 Fe, regulation toxicant concentration, 203 Fenton’s reagent, oxidant, 9 Fenvalerate, chemical hydrolysis, 23 Five-membered heteroarene oxidoreductase, 129 Flavanoids aerobic degradation, 560, 563 anaerobic degradation, 560, 562 anaerobic ring B reduction, 163 inducer PCB degradation, 210, 211, 464, 607 plant metabolite, PCB degradation, 607 Flavanone, Aspergillus transformation, 559 Fluoroacetate degradation, 381 metabolite, 381 Fluoranthene alternative bacterial dioxygenations, 405 quinones as bacterial metabolites, 406 transformation, Cunninghamella, 410 Fluorinated hydroxybenzoate decarboxylation, Candida, 19 F − NMR application, 287 Fluorobenzenes, toluene 2,3-dioxygenase transformation, 494 Fluorobenzoate, dioxygenation, 496 2-fluorobenzoate metabolism aerobic, 496 denitrifying, 499 3-fluorobenzoate, distal dioxygenation, 497, 498 4-fluorobenzoate 1,2-dioxygenation, 498, 499 hydrolytic fluoride loss, 498, 499 4-fluorobiphenyl, transformation, Tylospora, 495 Fluoromethane, oxidation, Nitrosomonas, 378 6-fluoro-3-methylphenol, methanogenic transformation, 502 1-fluoronaphthalene, metabolism, Cunninghamella, 495 5-fluoronicotinate, degradation, 534 Fluorophenol hydroxylation, 19 F − NMR application, 288 Fluoroquinolone ciprafloxacin, MnO2 transformation, 32 enrofloxacin, photolysis, 7 photolability, 539 Fluorotelomer alcohol, degradation, 382 Fluorotoluene, dehydrogenation to fluorobenzoate, 496 5-fluorouracil fungal transformation, 540 metabolism, 19F − NMR application, 287 Fluoxetine, (Prozac), abiotic degradation, 8 Formate dehydrogenase, 188 Formate interspecies transfer, 320 Free radical, toluene anaerobic activation, EPR application, 289 Fumarate anaerobic metabolic pathways alkane, 304, 305 cresol, 4502-methylnaphthalene, 409 toluene 390−391, 448 xylene, 681 Fungal reactions hydroxylation/ring fission, Paecilomyces, 77, 558 analogous to mammalian systems, 76 Furan-2-carboxaldehyde, anaerobic degradation, 553, 554 Furan-2-carboxylate, aerobic degradation, 523, 524 Furan-3-carboxylate, anaerobic degradation, 553, 554 Fusaric acid, bacterial degradation, 608

CRC_7241_Index.indd 700

Index

G Gallate (3,4,5-trihydroxybenzoate), anaerobic degradation, 450 Gentisate (2,5-dihydroxybenzoate) dioxygenase, 110, 125, 428, 430 pathway, 111, 425−426, 426 Geobacter Fe(III) electron acceptor, 153 reduction, 74 metallireducens, humic acid free radical production, EPR application, 289 sulfurreducens, oxygen as electron acceptor, 202 Glucose, enhancement/suppression of biodegradation, 198 Glutathione-mediated transformation, 307, 363, 369 Glutathione S-transferase, reactant, 307 Glycerol, anaerobic degradation, 310 Glycerol trinitrate reductase, 571 Glyoxal oxidase, 185 Phanerochaete, 133 Glyoxalase I, 182 Glyoxylate, aerobic metabolism, 317, 318 Glyphosate, degradation, 589 Gordonia, 65 Grass, naphthalene-[13C] mineralization analysis, 652

H Haloalkane degradation, 358−360 mycobacteria, 63 dehalogenase, structure, 290, 361 Haloalkanol dehalogenase, 361−362 2-Halobenzoate 1,2-dioxygenases, 425, 496 two component, 469 three component, 470 4-Halobenzoate, hydrolytic reactions, 472 Halogenated arene hydrocarbons, aerobic degradation, 455−458. See also PCB Halogenated benzoates. See also specific compounds degradation, Rhodococcus, 64 dioxygenation, 469−472 Halogenated phenols. See also specific compounds degradation, 482−486 hydroxylation, 483 Halogenated substrates, inhibition of degradation acyl chloride intermediate, 455 protoanemonin, 456 Halogenation, biotic, 134, 135 Haloperoxidase activities degradation of thiocarbamates, 116, 134 dihydrocoumarin hydrolase, 134−135 halogenation, 134, 135 Halophiles, 75 Heavy metal resistance plasmids, 224 Heme oxygenase, 114 Heptachlor transformations Daphnia magna, 99 photochemical, 6 Heteroarenes. See also specific compounds anaerobic degradation, 653 partial removal, microcosm, 647

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Index Hexachlorobenzene (HCB) anaerobic dechlorination/electron donor supplement, 663 partial anaerobic dechlorination, 489 reductive dechlorination, 458 Hexachlorobuta-1,3-diene, hydroxocobalamin dehalogenation, 27 Hexachlorocyclohexane (HCH) aerobic metabolism, 350−353, 354 isomers, 352 terminal metabolites, 352, 353 anaerobic transformation, 353−355 coupling to ATP synthesis, 355, 369 catabolic plasmid, 225 [36Cl] metabolism, 278 dechlorination, cyanobacteria, 61 degradation, acyl chloride intermediate, 122 enantiomeric biodegradation, α-isomer, 54, 353 extreme persistence, 662 hydroxocobalamin dehalogenation, 27 reductive dechlorination, Clostridium, 369 subcritical water degradation, 37 transformation, Citrobacter, 68 α-Hexachlorocyclohexane (HCH) aerobic mineralization, limitation by desorption, 209 Hexachloroethane anaerobic dehalogenation, 355 thiol reduction, 28 Hexadecane aerobic degradation at low oxygen concentration, 203 Hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX), degradation, 585, 676 Histidine aerobic degradation, 523, 524 ammonia lyase, 434 decarboxylation, 68 HMX degradation, 585 Humic acids bacterial reduction, 154 degradation role, 208 electron acceptor, 155 free radical production, EPR application, 289 photolytic role, aqueous media, 3 singlet oxygen (1O2) production, 12 Hybrid poplars, metabolism atrazine, 98, 603 creosote contamination, 652 hexahydro-1,3,5-trinitro-1,3,5-trizine (RDX), 603 trichloroethene, 98, 603, 685 Hydratase C=C,130 C≡C,131 R-C≡N,131 Hydrocarbons bacterial tolerance, 169 emulsifier synthesis/growth, 213 low temperature degradation, 200−201 Hydrofluorocarbon, monooxygenation, 379 Hydrogenases, 184, 188 nitroarene reduction, Clostridium, 509 Hydroperoxide, alkane aerobic degradation, 303 3-Hydroxyanthranilate, metabolite, 514 4-Hydroxybenzaldehyde rearrangement, 427

CRC_7241_Index.indd 701

701 Hydroxybenzoate, hydroxylation, 108 3-Hydroxybenzoate aerobic degradation, 428 fermentation, Sporotomaculum, 439 3-Hydroxybenzoate-6-hydroxylase, 108 4-Hydroxybenzoate anaerobic degradation, Thauera, 439 monooxygenation/NIH rearrangement, 426 oxidative decarboxylation, yeast, 427 reversible anaerobic decarboxylation, 431 4-Hydroxybenzoate-3-hydroxylase, 108 2-Hydroxybiphenyl metabolite dibenzothiophene, nitration, 57, 566−567 monooxygenase, 425 2-Hydroxyethylsulfonate, terminal electron acceptor, 590 Hydroxyl radicals, 4 Fenton´s reagent, 9 generation mechanisms, 4 toxicity to bacteria, 289 tropospheric formation, 14 3-Hydroxykynurenine pathway, tryptophan degradation, 526 4-Hydroxyphenylacetate hydroxylase, 111 monooxygenation/NIH rearrangement, 426 3-Hydroxyphenylpropionate, growth substrate, 69 4-Hydroxyphenylpyruvate dioxygenation, 126, 127, 426 Hydroxypurines, anaerobic hydroxylation, Peptococcus, 544 2-Hydroxypurine, dismutation Peptococcus, 544 4-Hydroxypyridine (pyridin-4-one), degradation, 530, 531 3-Hydroxyquinol-4-one dioxygenation, 539 2-Hydroxyquinoline-8-hydroxylase, 109 5-Hydroxysalicylate, ring fission, 425 Hydroxylamine-[15N reaction with humic/fulvic acid, NMR application, 287 Hydroxylase, 104. See also specific substrate fluoroacetate, 381 Hydroxylation/ring fission, Paecilomyces, 558 benzofuran, Paecilomyces, 558 biphenyl, hydroxylation/ring fission, 77, 558

I Iboprofen, degradation, 435 Immobilization, chlorophenol-degrading organisms, 659 Immobilized cells application to degradation, 261 chlorophenol-degrading organisms, 659 Indol-3-acetate, degradation, 526 Induction, catabolic enzymes, 195−198 analogue substrate, 195 induction by cosubstrates, 365 unrelated substrate, 197 Intradiol ring fission, 105, 110, 123 Fe2+ enzymes, 428 Intramolecular hydroxyl transfer, 1,2,3-trihydroxybenzene degradation, 450 isomerase, 450 Iodoacetate, product of chlorination (Cl2)/iodide, 32 Iodobenzoate. See 3,5-Diacetamino-2,4,6-triiodobenzoate Iron siderophore, 182

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702 Isoflavanoids, aerobic degradation, 560, 561 Isoflavanone, Aspergillus, transformation, 559 Isonicotinic acid hydrazide (INH), metabolic activation, 170 Isonitrile metabolism, 322 Isoprene, metabolism, 65, 307 Isoquinoline degradation, 539 oxidoreductase, 130, 164, 187, 539 Isotope fractionation factor, 281 ratio, 281 experimental determination, 282 Isotope δ13C fractionation, bioremediation, 629−632 Isoxaflutole, chemical hydrolysis, 24

K Karakin aglycone, electron acceptor, 586 Kinetic isotope effect, 281 3-Ketoadipate pathway, 110, 220−221, 428 2-Ketoglutarate anaerobic oxidoreductase, 164 dioxygenase, 125−126, 577 2,4-dichlorophenoxyacetate, 127 marker, 4-nitrophenoxyacetate substrate, 612 taurine, 590 Rhodotorula glutinis reactions, 129

L Lanosteroid, cytochrome P450 demethylation, 117 Limonene degradation, 339−340 transformation, by cutworm larvae of Spodoptera litura, 97 Linuron herbicide, adaptive degradation, 217 Lithotrophic bacteria, 60 Long-chain alkane degradation, yeast, 78 alkanoate, syntrophic degradation, 194 alkyl sulfonates, monooxygenation, 590 dialkyl ethers, degradation, 574 Low oxygen-concentration, degradation, 203−204

M Magnesium dioxygenase, 124 Malonate, anaerobic metabolism, 73, 318 Mandelate degradation, 433, 434 regulation, 221 Manganese dioxygenase, 124 Bacillus, Mössbauer spectroscopy, 290 enzymes, 181 Marine bacteria, degradation potential, 58 Marine ecosystems, biodegradation, 249 Marinobacter, alkane degradation, 58 Mecoprop [2-(4-chloro-2-methylphenoxy)propionate)], degradation, 54, 577 Media dilute complex, 58, 258 mineral basal, 252 solid, 257 Meisenheimer-type hydride complex, 164, 512, 516−517

CRC_7241_Index.indd 702

Index 2-mercaptobenzothiazole, degradation, 567 Mesocosms, 266 Metabolic markers dioxygenase enzymes 2-ketoglutarate, 612 naphthalene, 611 toluene, 611 metabolites arene dihydrodiols/PAH degradation, 611 benzylsuccinate/toluene, anaerobic degradation, 267 chlorobenzoates/PCB degradation, 664 methylbenzylsuccinates/xylene, anaerobic degradation, 681 7-Methylbenz[a]anthracene, transformation, Cunninghamella, 410 3-O-Methylgallate central metabolite, 429, 577 degradation, 429, 430 2-Methylnaphthalene, anaerobic metabolism, 408, 409 Metabolic association, 194 Metabolite conjugation in fish, 95 Metabolites/bioremediation, association with humic acid, 612 Metal/metalloid microbial resistance, 172−175 Metal cation/oxyanion reductase, 165 Metal cations, electron acceptor, 152−154 Metal concentration, metabolism, 253, 254−256 Metallothionein, 175 Metamitron degradation, 546 Methane aerobic degradation, 297 anaerobic oxidation, 304, 626 plant-root, bacterial degradation, 607 Methanearsonic acid, degradation, 594 Methane/propane/ammonia monooxygenase, chloroethene oxidation, 682 Methane monooxygenase, 103, 298 copper enzyme, 184, 298 versatility, 69, 70, 297−298 Methane sulfonate, monooxygenation, 590 Methanogenesis, 182, 183 Methanol, anaerobic oxidation, 310 Methanotrophs, 13C-methane analysis, 628 Methionine degradation/elimination, 579 salvage, enediol dioxygenation, 126, 128, 182 Methoxybenzoates aerobic degradation, 429 aerobic O demethylation, 577 anaerobic growth, 577 anaerobic O-demethylation, 577 sulfide-dependent, 578 Methylamine, degradation, 311 Methylamine-[15N] degradation, NMR application, 287 Methylation, metal/metalloid resistance mechanism, 173 Methylbenzylsuccinate markers, anaerobic xylene degradation, 681 Methyl chloride aerobic/anaerobic degradation, corrinoid pathway, 357 2-(2-methyl-4-chlorophenoxy)propionate (Mecoprop) enantiomeric preference biodegradation, 54, 577 removal, 672

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Index Methyl coenzyme M reductase, 182 sequence application, 626 structure, 291 Methyl fluoride, oxidation Nitrosomonas, 364 Methylglyoxal, degradation, 183, 310 N-methylisonicotinate, degradation, 532, 533 Methylmercury chloride, transformation to methane, 594 2-methylnaphthalene, anaerobic degradation, 409, 608−609 Methylotrophs, rice paddy, 13C-methanol analysis, 628 Methylparathion transformation, multiphase kinetics, 219 Methylphenylphosphonate, reduction to benzene, 67, 589 Methylquinolones, degradation, 538 Methyl tert-butyl ether (MTBE) bacterial metabolism, 686 degradation, 574 remediation, δ13C application, 629 TiO2 photochemical degradation, 10 transformation, 574 Metribuzin metabolites in leachate, 673 transformation, 546 Microcosm application, 264−265 13 C-substrate, DNA probe, 625 Mimosine, aerobic degradation, 532, 535 Mineralization, 247 Miserotoxin aglycone, electron acceptor, 586 Models of natural systems, 266 Molybdenum, oxidoreductase, 185−187 Molybdenum-dependent dehydrogenase, 185 W application, 280 Molybdopterin, 130, 185−187 Monochloroanilines, aerobic degradation, 490 Monofluorophenol, metabolism Rhodococcus, 500, 501 Monomethylsulfate, oxidative degradation, 570 Monooxygenase/dioxygenase, dual activity Mycobacterium, 406 Pseudomonas, 107, 400, 682–683 Monooxygenations, 106−113 alkene, 305 halogenated alkenes, 364 Monoterpenoid aerobic degradation, 300, 339−341 Morpholine degradation, 312 [13C] degradation, NMR, 286 cytochrome P450, 116 mycobacteria, 63 Mössbauer spectra, 57Fe application, 280 Mössbauer spectroscopy, application, 280, 290 Mutase, phenylhydroxylamine, 510 Mycobacteria degradation potential, 63 dehalogenase, 360 haloalkane degradation, 63 Mycobacterium chlorophenolicum detection by 16S rRNA / PCR, 659 Mycobacterium, dual monooxygenase/dioxygenase activity, 406 Mycorrhizosphere, Pinus sp., 609

N 15

N, NMR applications, 287 Naphthalene aerobic degradation, 399−400

CRC_7241_Index.indd 703

703 catabolic plasmid, 225 dioxygenation pathway, 399 pathway induction, 399 plasmid-borne 399 anaerobic mineralization, 408 fungal transformation, 411 metabolism, oxygenic phototrophs, 413 toxicity, low oxygen concentration, 204, 647 transformation to naphth-1-ol by Bacillus, 400 Naphthalene/2-methylnaphthalene, sulfate-reduction, enrichment factor, 630 Naphthalene carboxylates, aerobic degradation, 401 Naphthalene dioxygenase, 400 marker, indigo production, 611 rhizosphere gene, 652 Naphthalene monooxygense/dioxygenase, dual activity, 400 Naphthalene-2-sulfonate metabolites as redox mediators, 155 Naphthalene soil-sorbed, degradation, 210 Naphthalenesulfonates, degradation, 506−507 Naphthenates, degradation, 641 Naphth-1-oate, surfactant during growth, 213 Naphth-1-ol, degradation, 401 Natural consortia, 264 Natural substrates, induction of PCB metabolism, 664 Neptumonas, naphthalene dioxygenase, 59 Nickel enzymes, 182−184 Nicotinate anaerobic metabolism, 534−536, 536 hydroxylase Mo(V) EPR spectrum 280 77 Selenium, application, 280 Nicotine aerobic degradation, 132, 532, 533 diazodiphenoquinone, chemical artefact, 532 catabolic plasmid, 225 oxidase, 132 18 O application, 279 NIH shift − metabolism, 426 arene oxide bacteria, 107 fungal transformation, 410 2-chlorobiphenyl hydroxylation, 464 1,3-dichlorobenzene monooxygenation, 458 2,4-dichlorophenoxyacetate fungus, 486 4-hydroxybenzoate degradation, 109 4-hydroxyphenyolacetate degradation, 426 Nitrate anaerobic electron acceptor, 148 PAH degradation, 647 assimilation, 148 Nitrate ester reductase, 162 reductive degradation, 571 Nitrate radicals, 14 Nitrated metabolites from nitrite, 57 Nitrilase, selectivity, 322 Nitrite, metabolite incorporation, 57 Nitroalkane degradation, Fusarium oxidase, 133, 586 electron acceptor, 156 oxidation, 129

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704 Nitroarenes. See also specific compounds bioremediation, problem areas, 675, 677 degradation, mycobacteria, 63 partial reduction to hydroxylamine, 510 reductase, 162 three component dioxygenase, 512 Nitrobenzene degradation dioxygenation, 512 reduction/rearrangement, 511 2-nitrobenzoate, reductive metabolism, 515 anthranilate pathway, 514 hydroxylamine/rearrangement, 514 3-nitrobenzoate, dioxygenation with nitrite loss, 514, 515 4-nitrobenzoate, reduction to hydroxylamine/ rearrangement, 514, 515 Nitrodiphenylamine, nitro group reduction, 510 Nitrofluoranthenes, atmospheric production, 21 2-nitronaphthalene, atmospheric production, 20 Nitrophenols, secondary atmospheric oxidation products from monocyclic arenes, 21 2-nitrophenol, degradation by monooxygenation, 515 3-nitrophenol degradation by reduction/rearrangement, 516 4-nitrophenol adaptive degradation, 217 degradation, 515−516, 516 monooxygenation, 112 2-nitropropane dioxygenation, yeast, 586 oxidation, 128 3-nitropropanol, electron acceptor, Denitrobacterium, 586 Nitropyrenes, atmospheric production, 21 1-nitropyrene, phototransformation, 11 N-nitrosodimethylamine monooxygenation, 586 Nitrosomonas europaea oxidation, 70−71 haloalkane oxidation, 364 2-nitrotoluene dioxygenase, three component, 512 4-nitrotoluene oxidation to nitrobenzoate, 513 3-nitrotyrosine, degradation, 515 NMR technical developments, 284 NMR, application 13 C, 285 19 F, 287 15 N, 287 17 O, 287 31 P, 288 Nonafluoropentanoic acid, polyoxometalate degradation, 10 Non-oxidative decarboxylation, 430−431 Nonsymbiotic diazotroph, 608 4n-nonylphenol metabolism, plant cultures, 605 Nonylphenol, degradation, 301, 302, 448 side-chain rearrangement, 301−302, 427, 448 Nonylphenol, polyethoxylate partial degradation, 575 NTA degradation, 312 Nutrient media, metabolic effects, 219−220

O Octachlorodibenzo[1,4]dioxin anaerobic partial dechlorination, 467, 666 degradation, Phanerochaete, 667 photochemical product from pentachlorophenol, 7

CRC_7241_Index.indd 704

Index Octachlorostyrene, blue mussel transformation, 96 Octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine (HMX), degradation, 585, 676 Octane, cytochrome P450 hydroxylation, 114 Octyl-2-sulfate, hydrolysis, 570 O-demethylation aerobic, 577 anaerobic, 577 sulfide-dependent, 578 Old yellow enzyme reaction α,β-unsaturated ketone, 339 reductase, 163 reduction α,β-unsaturated nitro compounds, 586 Oligotrophs, 211 Oligotrophic bacteria, 59 On-site vs off-site PAH remediation, 645 Organophorphorus pesticide hydrolysis, 677 Organophosphate as source of phosphorus, 571 Organophosphorous acid anhydrase (OPA), 571 Orotic acid, anaerobic degradation, 541 o-phthalate. See also Phthalate dioxygenation, 532 metabolite from degradation anthracene, phenanthrene, 404 isoquinoline, 539 Oxaarenes cytochrome P450 transformation, 553, 554 three component 4,4a-dioxygenase, 555, 556 Oxalate anaerobic metabolism, 73, 318 decarboxylase, 185 Oxidoreductases, 129−130. See also Pyridine, Isoquinoline, Quinoline Oxygen concentration as determinant, 201 synthesis of toxicant, 203 tolerance Clostridium during microaerophilic growth, 202 Desulfovibrio vulgaris, 202 Oxygenases, 104, 105 Oxygenic phototrophs, metabolism, 413

P PAH degradation. See also specific compounds adaptive, 216 aerobic bacteria, 403 aging, 208 cis dihydrodiol markers, 611 combined biological/chemical treatment, 645 fractionation factor, 630 growth with phenanthrene as substrate, 196 induction, 221 Leclercia, 69 mycobacteria, 63 PAH degradation (contd.) range of bacteria, 647 subcritical water, 37, 645 stereospecificity, 403 white-rot fungi, 77 PAH metabolism anaerobic, 652 cytochrome P450 hydroxylations, Streptomyces, 115

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Index fish, 92−93 fungal transformations 76, 409−413, 646 H2O2 requirement, 650 mammalian metabolism model, 409 phenolic metabolites, 410 quinone production, 651 stereospecificity compared, 409 fungus-mediated polymerization, 651 ozone oxidation, 31 polychaetes, 97 sorption/desorption, 208, 646 surfactants, 649−650 transformation, phototrophic bacteria, 61 PAH quinones, abiotic formation, 653 PAH remediation combined biological/chemical treatment, 645 conclusions, 654 gas works site, application of fungi, 646 naphthalene dioxygenase genes, 623 oxygenase genes, PCR amplification, 623 plant role, 623 prairie grass stimulation, 605 primary targets, 644 white-rot fungi, 651 Paraquat photolysis, 8, 532 Partition/organic matrix, contaminants, 601 PCB/PBB, adaptive degradation, 217 PCB atmospheric hydroxylation, 16 bioavailability, surfactants, 665 chlorobenzenes as degradation markers, 267, 611 remediation, 665−666 ring fission, 461 sediment association, 209 sub-critical water dechlorinations, 25, 37 PCB aerobic degradation, 459−464, 462, 664 chloroacetophenone metabolites, 463 flavanoids as inducers, 210, 464, 607 hydrolytic reactions after ring fission, 461−462 long-term exposure, 665 organisms, 459 rhizobia, 67 tree proximity, 665 PCB anaerobic dechlorination, 465−467 Chloroflexi group, 466 Dehalococcoides group, 466 metabolic inhibitors, 626 non-methanogenic mixed culture, 466 patterns, N-, P -, LP -, 465, 665−666 preferential congener removal, 666 temperature significance, 664 δ13C fractionation, 629 PCB metabolism fish hydroxylation, 95 Nereis virens, 97 Phanerochaete chrysosporium, 77, 465 plant cultures, 604 Pentachloroaniline, partial anaerobic dechlorination, 490 Pentachloroanisole, Phanerochaete formation/ mineralization, 659

CRC_7241_Index.indd 705

705 Pentachlorobenzene hydroxylation, P450 cam mutants, 458 partial anaerobic dechlorination/electron donor, 663 Pentachloronitrobenzene transformation methanogenic, 673 Tetrahymena thermophila, 99 Pentachlorophenol abiotic degradation, glow discharge plasma, 38 association with soil organic matter, 206 biodegradation, 485, 486 acyl chloride intermediate, 122 additional substrate requirement, 659 Fenton´s reagent transformation, 9 monooxygenation, 112 dehalogenation, 486 photochemical transformation, 5 octachlorodibenzo[1,4]dioxin, formation, 7 stream degradation, 266 tolerance to aerobic bacteria, 659 Pentachlorophenol metabolism Phanerochaete chrysosporium, 76 Eichornia crassipes, 97 plant, 604 Pentachloropyridine, partial anaerobic dechlorination, 534 Pentaerythritol tetranitrate degradation, 69, 571 reductase, 571 Pentafluorophenol, monooxygenation, 500 Perfluoroalkanecarboxylates degradation partial photochemical, 667 sulfate radical, 12 Perfluorooctanoate, photochemical degradation, 12 Peroxyalkanoyl nitrates, photochemical formation, 13 pH buffer control, 254 Phenanthrene anaerobic mineralization, 408 bacterial aerobic degradation, 403, 404 degradation, humic matter-associated, 209 dioxygenase, 403 growth substrate for PAH degradation, 196 metabolism by oxygenic phototrophs, 413 transformation Phanerochaete, 415 Streptomyces to trans dihydrodiol, 405 Phenazine, 2-aminodiphenylamine metabolite, 510 Phenazine-1-carboxamide, bacterial metabolite, 607 tomato rot control, 608 Phenazine-1-carboxylate, bacterial metabolite, 607 Phenmedipham, chemical hydrolysis, 23 Phenol degradation anaerobic, 449−452 dehydroxylation, 165, 187 Aspergilllus fumigatus, 77 community, 13C-phenol analysis, 627−628 strain introduction, 624 Variovorax sp., 625 Phenol metabolism conjugation by fungi, 410 hydroxylase, 446 gene, 624 monooxygenation, 110, 446 yeast, 77−78

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706 Phenol-degrading community, 13C-phenol analysis, 627−628 Phenoxyalkanoate degradation, 576 dioxygenase dealkylation, 577 enantiomeric specificity, 577 Phenylacetate degradation, Escherichia coli, 435 Phenylacetate-[13C] aerobic degradation, NMR application, 286 l-phenylalanine aerobic transformation, 113 ammonia lyase, 434 anaerobic degradation, Thauera, 436 hydroxylation, 113, 426 3-phenylbutyrate, degradation, 435 enantiomeric specificity, 54, 435 4-phenylcatechol, toxic metabolite, 222 Phenylhydroxylamine, rearrangement, 510 Phenylmercuric acetate, transformation to benzene, 594 Phenylphosphate synthesis, anaerobic degradation of phenol, 449 Phenylurea herbicide-degrading population diversity, 625−626 Variovorax sp., 616, 625 Phloroglucinol (1,3,5-trihydroxybenzene) degradation, Fusarium, 449 metabolite from 3,4,5-trihydroxybenzoate, 450 Phosphatase, 571 Phosphonatase, 588 Phosphonates degradation pathway, 588−589 phosphorus source, 588 Phosphonomycin, degradation, 67 Phosphorofluoridates, degradation, 677 Photochemical experimental generation of reactants, 245−246 Phototrophic bacteria, transformation, 61 Phoxin transformation by Glycine max, 97 Phthalate. See also o-phthalate aerobic degradation, 431 dioxygenase, 431 esters, hydrolysis, 569 Phthalazine, hydroxylation, 539 Phytoremediation, 602−603 α-pinene, degradation, 341 Plant cultures, metabolism 4n-nonylphenol, 605 pentachlorophenol, 604 PCB, 604 Plant cuticles, roles in photochemical transformations, 11−12 Plant exudates, 605 metabolism, 97−98 metabolites, PCB degradation, 665 remediation primary role, 603 secondary role, 605 Plasmids catabolic, 224−227, 225 transmission, 226 Poly(cis-1,4,isoprene), aerobic degradation, 66, 305 Polybrominated biphenyls (PBB)

CRC_7241_Index.indd 706

Index anaerobic debromination, 467, 468, 666 primers, dechlorination PCBs, 467 Polychlorinated dibenzo[1,4]dioxin, atmospheric hydroxylation, 16 Polyethylene glycol degradation aerobic, 67, 575 anaerobic, 575 Polyfluorinated phenol, hydroxylation, Trichosporon, 500 Polyfluorinated substrates, metabolism, 382 Polyhalogenated alkanes, cytochrome P450 reductive dehalogenation, 366 Polyisoprenoid oxygenase, 128 Polyoxometalates (heteropolyacids) degradation 2,4,6-Trichlorophenol, 10 oxidants, 10 Porphyrin-mediated dehalogenation, 26−27 Porphyrobacter, biphenyl and dibenzofuran degradation, 59 Prairie grass-stimulated PAH remediation, 605 Prednisone reduction, “old yellow enzyme”, 163 Pre-exposure enhanced degradation, 601 naphthalene degradation, 216 Pregnenolone, C17 side-chain removal by cytochrome P450, 117 Primers for anaerobic dehalogenation PCB, 465, 467 octachlorodibenzo[1,4]dioxin, 467 Pristane, aerobic degradation, 300 Pristine environments degradation, 215−216 degradative bacteria, 622 Prodrug metabolism, 170 Progesterone, cytochrome P450 15β-hydroxylation, 115 Propane exceptional oxidation by Rhodococcus, 126, 303 Propan-2-yl sulfate, oxidative degradation, 570 Propham, abiotic degradation, 3 Propionate aerobic degradation, 299 anaerobic metabolism, 319 [13C], NMR application, 286 analysis, [13C] syntrophic oxidation, 628 syntrophic, 194, 320 Protoanemonin 4-chlorobenzoate synthesis, inhibition of PCB degradation, 223 metabolic inhibition, 463 Pseudomonad monooxygenase/dioxygenase, chloroethene oxidation, 682−683 Psychrophiles, 74 Purine aerobic degradation, 542 anaerobic degradation, 544−545, 545 fermentation, clostridia, 73 hydroxylase, 130, 545 Pyrene aerobic degradation, 406−407,407 chlorination (Cl2), 33 degradation at low oxygen concentration, 203 dihydrodiol marker for aerobic degradation, 267 hydroxylation by Porcellio scaber, 97 mineralization, [13C] pristine forest soil, 646 oxidation by Nematoloma frowardii, 77

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Index transformation Cunninghamella, 412 white-rot fungi, 415−416 Pyrene quinone bacterial metabolite, 407 white-rot fungus metabolites, 416 Pyridine aerobic degradation, 529 dioxygenation, 532, 534 oxidoreductase, 129 ring reductive pathway, 529 Pyridine carboxylates degradation, 531 interacting pathways, 530 Pyridine-4-carboxylate, degradation, 530 Pyridine dicarboxylates coupled photolytic/microbial degradation, 13 oxidation, 532 Pyridine-2,6-dithiocarboxylate, function in tetrachloromethane degradation, 363 Pyridoxal, degradation, 529 Pyrimidines. See also specific compounds aerobic/anaerobic ring reduction, 540, 541 degradation, clostridia, 73 Pyrrole-2-carboxylate, aerobic degradation, 523, 524 Pyruvate, anaerobic metabolism Pseudomonas, 201 Pyruvate formate lyase mechanism, EPR application, 289 Pyruvoyl enzyme, amino acid metabolism, 315

Q Quercitin anaerobic degradation, 560, 562 dioxygenase, 126, 185 Quinazoline, hydroxylation, 539 Quinoid mediators, azoreductase, 521 Quinoline aerobic degradation, 536−539, 538 dioxygenation, 537 hydroxylation, 536 oxidoreductase, 130, 537 Quinolone. See also Methylquinolone degradation, 536 pathway, tryptophan degradation, 526 Quinol-2-one 5,6-dioxygenase, single component enzyme, 537 Quinol-4-one-2-carboxylate, dioxygenation, 537 Quinone-mediated redox reductions azo dyes, 155, 521 nitroarenes, 28−29 Quinones bacterial metabolites fluoranthene, 406 pyrene, 407 pyrene and benzo[a]pyrene, Cunninghamella, 412

R Raleigh model, 281 RDX/HMX, subcritical water degradation, 36 RDX degradation, 585 partial metabolism

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707 hybrid poplar, 98 Reed Canary grass, 99 partial reduction, Klebsiella, 69 subcritical water degradation, 37 Reaction rates, kinetic aspects, 218−219 Recalcitrance, biodegradation, 247 Redox potential determinant, 205 growth media, 256 Redox mediators cinnaquinone, secondary metabolite from Streptomyces, 28 naphthalene-2-sulfonate metabolites from Sphingomonas, 155, 521 Regulation, degradative pathways, 220−222, 413 Remediation, chemical procedures, 602 Resorcinol, aerobic degradation, 448 Reverse methanogenesis, anaerobic methane oxidation, 304, 626 Rhizobia degradation potential, 66 PCB degradation, 67 Rhizosphere alkane monooxygenase gene, 652 catechol 2,3-dioxygenase gene, 652 colonization, phase variation in bacteria, 608 Rhodamine B, polyoxometalate degradation, 10 Rhodococcus metabolic versatility, 64−65 Riboflavin, degradation, 542, 543 Rimsulfuron, chemical hydrolysis, 23 Ring fission dioxygenases, 123, 428–429, 428, 430 Roseovarius, dimethylsulfoniopropionate degradation, 59 Rutin dioxygenation Aspergillus, 559 carbon monoxide metabolite, 128, 129

S Salicylate degradation, 425−426 5-hydroxylase, 108, 109 Salinity evaporation pond, population, 627 Secondary metabolites produced by bacteria, 607 Selenate electron acceptor, 151 reductase, 151 Sequential microbial/chemical reactions, 55 Selenium requirement purine hydroxylase, 545 xanthine dehydrogenase, 544 Silicibacter, dimethylsulfoniopropionate degradation, 59, 580 Simulated tropospheric reactions, 16, 245 Singlet oxygen (1O2), reactant, 12 Soil slurries vs solid phase microcosm system, 647 Solid media, preparation, 257 Sphingomonad classification, 67 Squalene, aerobic degradation, 305 Stable enrichment cultures, 260 Steroid 3-keto-4,6-diene, anaerobic reduction, 163 Steroids cytochrome P450 demethylation, 117

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708 Steroids (contd.) degradation, 342, 343 fungal hydroxylation, 116, 341 Stilbene, oxidative fission, 392 Storage of samples, 247 Streptomyces, metabolic potential, 65−66 PAH hydroxylation, 115 phenanthrene trans-9,10-dihydrodiol, 405 Structure-activity relationship, reaction rates, 219 Styrene metabolism ring dioxygenation, 392 side-chain monooxygenation, 392 Subcritical water, 25, 36−37 degradation hexachlorocyclohexane, 37 PAH, 644 PCB, 37 RDX/HMX, 36 Substrate concentration degradation, 248 desorption as limiting factor, 208, 209, 210 influence on degradation, 211−212 particulate matter attachment, 265 remediation significance, 612−613 Succinyloxyacetate degradation, lyase, 574 Sulcotrione, photochemical transformation, 12 Sulfamethoxazone, chlorination (Cl2), 33 Sulfate radical, oxidant, 32 Sulfate-reducing anaerobic bacteria, 73 Sulfides, organic enantiomeric oxidation, 580−581 oxidation, Nitrosomonas europaea, 71 Sulfoacetaldehyde central metabolite from C2 sulfonates, 590 degradation, 590 Sulfometron, chemical hydrolysis, 23 Sulfonates bacterial rhizosphere survival, 605 electron acceptor, 590 Sulfone metabolites, aquatic mammals, 96 Sulfonylurea, cytochrome P450 hydroxylation, 65, 116 Sulfosuccinate, degradation, 590 sulfur sources, phototrophic bacteria, 61 Superoxide dismutases, 182, 303 Surfactant, naphtho-1-ate growth, 213 Syntrophic degradation, 194

T Taurine degradation, 590 electron source, Rhodopseudomonas, 590 TCA cycle modified, acetate anaerobic metabolism, 319 Temperature anaerobic dehalogenation, 466−467 bioremediation, 613 degradation determinant, 200 Terminal metabolites 2-aminophenol, 2-aminophenoxazine-3-one, 510 bioremediation, adverse prognosis, 612 hexachlorocyclohexane, 352, 353 nitrated metabolites from nitrite, 57 trifluorobenzoates, 503

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Index Terpenoid, PCB degradation, 210, 464, 607 Tetrabromobisphenol-A, anaerobic debromination, 489 Tetrachloroaniline, partial anaerobic dechlorination, 489 1,2,3,4-tetrachlorobenzene, degradation, 456 1,2,4,5-tetrachlorobenzene, degradation, 456,503 Tetrachlorobiphenyl congeners, TiO2 photochemical degradation, 10 2,3,4,6-tetrachlorobiphenyl dechlorination, temperature effect, 200 1,2,3,4-tetrachlorodibenzo[1,4]dioxin, CaO transformation, 29 2,3,7,8-tetrachlorodibenzo[1,4]dioxin degradation, Phanerochaete, 667 anaerobic dechlorination in aged sediment, 667 Tetrachloroethene anaerobic dechlorination, 366−369, 367 Clostridium bifermentans, 368 Desulfurospirillum, 367 Methanosarcina, 368 strain KB-1, 685 Sulfurospirillum multivorans, 74, 684 37 Cl fractionation factor, 632 dehalogenase, 684 electron acceptor, Desulfitobacterium metallireducens, 74 partial reduction, Enterobacter, 69 reactor dechlorination, 685 reentry of tropospheric transformation products, 19 Tetrachloromethane bacterial supplementation/electron donor, 682 degradation, 149 EPR application, 289 pyridine-2,6-dithiocarboxylate, 289, 363 [13C] degradation, 277, 289 partial anaerobic dechlorination, 370 thiol reduction, 28 Tetraethyl lead, degradation, 593 1,1,1,2-tetrafluoroethane, reentry of tropospheric transformation products, 19 Tetrahydrofuran, growth substrate, 576 Tetrahydronaphthalene (tetralin), degradation, 337, 338 Thermophiles, 75 Thiocarbamate, metabolism, 322−323 Thiocyanate, metabolism, 324 Thiol reduction, 27 Thiophene-2-carboxylate, aerobic degradation, 523, 524 Threshold concentration, degradation, 212 Thymidylate synthetase, gratuitous dehalogenation, 541 Thymine, aerobic degradation, 540 TiO2 surface, photochemical degradation, 10 α-tocopherol, nitrated metabolite from nitrite, 57 Toluene as metabolite phenylalanine, anaerobic, 317 TNT − nitrogen source, 511 Toluene/phenol enzyme induction for trichloroethene degradation, 197 Toluene/phenol oxygenase induction for chloroethene remediation, 682−683, 683 Toluene 2,3-dioxygenase marker, indigo production, 611 Toluene activation to benzyl radical, 390−391 EPR application, 289 aerobic degradation, 387−388

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Index chromosomal, 387 methyl oxidation, 387 ring dioxygenation, 387 amino acid anaerobic metabolite, 317 anaerobic degradation, 389−391 [13C], 630 benzylsuccinate marker, 267 condensation reactions, 390 methyl group dehydrogenation, 389 degradation desorption as limiting factor, 209 low oxygen concentration, 203 2,3-dioxygenase, 387 monooxygenation, 107−108, 387−388 plasmid-borne degradation, 387 plasmid/chromosomal genes, 226 Toluene-4-sulfonate degradation, 506 transport, 214 Toxaphene, partial anaerobic dechlorination, 370, 662 Toxic metabolites, 222−224 Toxic substrates, experimental procedures, 256 Transhalogenation, haloalkanol dehalogenase, 362 Transport mechanisms, 214−215 systems benzoate, 214 EDTA, 215 4-chlorobenzoate, 214 2,4-dichlorophenoxyacetate, 214, 215 toluene-4-sulfonate, 214 1,2,4-triazine degradation, 546 1,3,5-triazine, nitrogen/sulfur source, 545 2,4,6-tribromophenol, anaerobic debromination, Desulfovibrio, 488 2,3,5-tribromopyrrole, product from chlorination with (Cl2)/bromide, 331 Tributyl tin aerobic transformation, 593 degradation, 3 2,2,2-trichloroacetaldehyde, dismutation, 364 Trichloroacetate aerobic degradation, 362 anaerobic transformation, Geobacter, 362 1,1,2-trichloro-1,2,2-trifluoroethane, reductive dechlorination, 379−380 1,2,3-trichlorobenzene bioremediation, methane monooxygenase, 663 1,2,4-trichlorobenzene, degradation, 456, 457 concentration dependence, 219 2,3,5-trichlorobenzoate, anaerobic dechlorination, 475 1,2,3-trichlorodibenzo[1,4]dioxin, dioxygenation, 558 1,1,1-trichloroethane photochemical transformation, 6 reentry of tropospheric products, 19 Trichloroethene metabolism aerobic degradation isotope fractionation, 631 toluene/phenol induction, 197 anaerobic dechlorination, δ13C, 631 electron donors, 685 ethene production, 685

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709 dehalogenase, 684 non-established dechlorinating taxa, 685 metabolism, hybrid poplar, 98, 603 methane/ammonia-oxidizing bacteria, 624 monooxygenation, 364, 365 population analysis, 624 removal hybrid poplar, 685 stimulated degradation, grass, 605 TiO2 photochemical degradation, 10 toluene-2-monoxygenase probe, 624 zero-valent iron transformation, 26 Trichlorofluoromethane cytochrome P450 degradation, 379 reductive dechlorination, Methanosarcina, 379 2,4,5-trichlorophenol degradation, 484, 485monooxygenation, 112 2,4,6-trichlorophenol degradation, 484, 485 polyoxometalate, 10 population, analysis, 625 pristine environment, 216 monooxygenation, 112 1,2,3-trichloropropane, aerobic degradation, 365 3,5,6-trichloropyridine-2-ol, aerobic degradation, 534 Triclopyr, aqueous photolysis, 3 Triclosan, MnO2 transformation, 32 Trifluoroacetate, potential atmospheric photochemical product, 19 Trifluoroethene, methane monooxygenase oxidation, 378, 381 Trifluoromethylarenes, resistance to biodegradation, 667 Trifluoromethyl benzoates cometabolism, 503 dioxygenation, 502−503 terminal metabolites, 503 3-trifluoromethylbenzoate, coupled photolytic/microbial degradation, 13 3-trifluoromethyl-4-nitrophenol, photochemical transformation, 5 2-trifluoromethylphenol transformation, Bacillus, 503 Trifluralin, photochemical degradation, 6 Trihalopropanes, degradation, 362, 365 1,2,3-trihydroxybenzene (pyrogallol), anaerobic degradation, 450 1,2,4-trihydroxybenzene dioxygenase, Phanerochaete, 129 metabolite from 2,4-dichloro- and 2,4-dinitrophenol, Phanerochaete, 129 1,3,5-trihydroxybenzene (phloroglucinol) degradation, Fusarium solani, 77 2,4,6-/3,4,5-trihydroxybenzoate, anaerobic degradation, 439 2,4,6-trihydroxypurine (uric acid) aerobic degradation, 542, 543 anaerobic degradation, 542 2,4,6-trinitrocyclohexanone, 516 2,4,6-trinitrophenol aerobic degradation, 65, 164 transformation, 516 2,4,6-trinitrotoluene (TNT) aquatic plant metabolism, 604 nitrogen source, formation of toluene, 511 reduction

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710 2,4,6-trinitrotoluene (TNT) (contd.) additional fermentable substrates, 675 amine polymerization, 675 Clostridium, 509 Myriophyllum spicatum, 98 Triphenyltin, pyoverdine transformation to benzene, 593 Tropic acid, degradation, 434 Tropospheric reactions, 14−15 Tryptophan degradation, 524−526 3-hydroxykynurenine pathway, 526 quinolone pathway, 526 Tryptophan dioxygenase, 124 Tryptophanase, 525 Tungsten enzymes, 187−188 Tyrosine, degradation, 426

U Uracil, degradation aerobic, 540 anaerobic, 541 α,β-unsaturated nitro compound, reduction, 586 Urease, 182 Urocanase, 523

Index Volatile substrates, experimental procedures, 257 Volatilization, contaminants, 601

W Water solubility, determinant, 213 White-rot fungi bioremediation, 622 degradation, 76−77 oxidation systems hydroxylase, 414 monooxygenase, 414 peroxidase, 414

X Xanthine dehydrogenase, 130, 544 Xanthine, central anaerobic metabolite, 544 Xanthone, aerobic degradation, 555, 556 Xylene/cresol, anaerobic degradation, enrichment factors, 630 Xylene aerobic degradation, 388−389 anaerobic degradation methylbenzylsuccinate markers, 681 sulfate-reduction, 391 Xylenols, aerobic degradation, 446

V Vanadium enzymes, 188 Vanillate, aerobic degradation, 433 Vanillyl alcohol degradation, oxidase, 133 Verrumicrobia, 58 Vibrionaceae, degradation, 69 Vinyl chloride aerobic degradation, 307 epoxidation, 364−365 metabolite, anaerobic dechlorination, 366, 367 microbial hydrolysis, 364 recalcitrance, 684 Vitamin additives, 254

CRC_7241_Index.indd 710

Y Yeasts alkane, degradation, 78, 303 aromatic substrate degradation, 77−78 biphenyl, hydroxylation/ring fission, Paecilomyces, 77, 558 ring fission, 5-chloropyrogallol, Candida, 487

Z Zero-valent iron reduction, 25

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